CHROMIUM TRANSFORMATIONS DURING AGING IN CONTAMINATED SOILS By Jingjing Shi B.Sc., Wenzhou University, 2013 M.Sc., University of Northern British Columbia, 2016 DISSERTATION SUBMITTED IN PARTIAL FULFILLMENT OF THE REQUIREMENTS FOR THE DEGREE OF DOCTOR OF PHILOSOPHY IN NATURAL RESOURCES AND ENVIRONMENTAL STUDIES UNIVERSITY OF NORTHERN BRITISH COLUMBIA April 2022 © Jingjing Shi, 2022 i ABSTRACT Extensive industrial activities have led to Cr contamination in the environment, which poses threats to ecosystems and human health. Currently, very little is known regarding the Cr accumulation in soils during aging. The purpose of this dissertation was (i) to enhance our understanding of Cr transformations occurring over a long period of time; (ii) to investigate the reduction of toxic Cr(VI) to non-toxic Cr(III) by soil microorganisms for developing bioremediation strategies for Cr-contaminated sites. Soil samples were collected from a longterm (> 30 years) tannery waste contaminated area in Shuitou (China) for this research. Chemical extraction methods showed the Cr(III) form was dominant (> 96.7% of total Cr) in these aged Cr-contaminated soils, with toxic Cr(VI) up to 144 mg kg-1. Of the total Cr(VI) present, immobile Cr(VI) represented > 90%. Synchrotron-based X-ray near edge structure spectroscopy demonstrated Cr species present were CrFeO3, CrOOH, and CaCrO4. The occurrence of immobile Cr(VI) species in long-term contaminated soils was further verified by a spiking experiment over 240-day aerobic incubation. Available Cr(VI) in soils continually decreased during aging, with immobile Cr(VI) increasing by 4.5 – 31% and immobile Cr(III) increasing by 68 – 95% of total spiked Cr(VI). These findings reveal that Cr(VI) reduction and immobilization were occurring concurrently in soils. Cr(VI) reduction occurs in soils with low pH and high organic carbon content via both chemical and biological processes, while Cr(VI) immobilization occurs in soils with cations (such as Ca2+) and Fe oxides. Shotgun metagenomic sequencing was used to analyze the microbial community composition in the soils and a batch solution experiment was employed to determine the Cr(VI) reduction capacity by soil microbial consortia. The results demonstrated the accumulation of high levels of Cr in a soil (e.g., 3141 mg kg-1 in S3-2) led to the increased abundance of Cr resistant and reducing microorganisms: ii Proteobacteria (69.9%) at phylum level, Betaproteobacteria (39.1%) at class level, and Massilia (12.6%) and Bacilli (0.57%) at genus level. Batch experiment results showed the addition of 1.0 g Cr-contaminated soils reduced 10 – 20 mg L-1 Cr(VI) in 20 mL of K2Cr2O7 solution at the condition of 30 oC at pH 7.8 – 8.0 within 7 days anaerobically and aerobically, when supplied with 0.2 g L-1 of Na-acetate as carbon and electron sources. The amount of Cr(VI) removed was highest (29.0 mg L-1) at 40 mg Cr(VI) L-1 in the presence of soil S3-2. Therefore, prospective application of mixed microbial consortia from high Cr-contaminated soils for bioremediation of Cr(VI)-polluted environments could be expected. iii CO-AUTHORSHIP I was the principal investigator of this dissertation including design of studies, acquisition of data, analysis of data and writing all the chapters. Chapters 3, 4 and 5 are each based on manuscripts. As the first author of these papers, I wrote the manuscripts and was responsible for incorporating comments and feedbacks in the revised manuscripts. Dr. Bill McGill supervised the experiments and contributed to experimental design, data analysis, and revision of the manuscripts. Dr. Todd Whitcombe and Dr. Mike Rutherford contributed to the design and implementation of experiments and helped review and improve the manuscripts. They were all included in authorship on all resulting publications. Dr. Ning Chen helped with the synchrotronbased XANES experiments and Dr. Wei Zhang provided useful suggestions on aging experiment, so they were included in related publications. Publication and authorships from this dissertation (Published or prepared for submission): Shi, J., McGill, W. B., Chen, N., Rutherford, P. M., Whitcombe, T. W., & Zhang, W. (2020). Formation and immobilization of Cr (VI) species in long-term tannery waste contaminated soils. Environmental Science & Technology, 54(12), 7226-7235. DOI: 10.1021/acs.est.0c00156. (Chapter 3) Shi, J., McGill, W. B., Rutherford, P. M., Whitcombe, T. W., & Zhang, W. (2022). Aging shapes Cr (VI) speciation in five different soils. Science of The Total Environment, 804, 150066. DOI: 10.1016/j.scitotenv.2021.150066. (Chapter 4) Shi, J., McGill, W. B., Rutherford, P. M., & Whitcombe, T. W. (2022). Metagenomics reveals microbial community in chromium contaminated soils and its direct use for bioremediation. Submitted to Journal of Hazardous Materials on April 1st, 2022. (Chapter 5) iv TABLE OF CONTENTS ABSTRACT ii CO-AUTHORSHIP iv TABLE OF CONTENTS v LIST OF TABLES ix LIST OF FIGURES xi LIST OF ABBREVIATIONS xiv ACKNOWLEDGEMENT xvi Chapter 1 GENERAL INTRODUCTION 1 Chapter 2 LITERATURE REVIEW 6 Abstract 6 2.1 Introduction 7 2.2 Occurrence of chromium 8 2.2.1 Naturally geogenic Cr 8 2.2.2 Anthropogenic Cr 9 2.2.3 Regulated levels worldwide 12 2.2.4 Reported contaminated sites worldwide 12 2.3 The fate of chromium in soils 14 2.3.1 Speciation 14 2.3.1.1 Oxidation states 14 2.3.1.2 Environmental availability 16 2.3.2 Aging processes 18 2.4 Environmental impacts 18 2.4.1 Toxicity of Cr(VI) and Cr(III) 18 2.4.2 Toxicity to the environment 19 2.5 Bioremediation of Cr-contaminated sites 20 2.5.1 Remediation strategy by reduction 20 v 2.5.2 Microbial resistance 24 2.5.3 Microbial Cr(VI) reduction 26 2.5.4 Factors affecting microbial Cr(VI) reduction 30 2.5.5 Pure cultures or mixed bacterial consortia 33 2.5.6 Biostimulation or bioaugmentation 34 2.5.7 The fate of Cr(III) 34 2.6 Summary 35 Chapter 3 FORMATION AND GENESIS OF IMMOBILE CR(VI) IN LONG-TERM CONTAMINATED SOILS 37 Abstract 37 Graphical Abstract 38 3.1 Introduction 39 3.2 Materials and methods 40 3.2.1 Study area and soil sample collection 40 3.2.2 Characterization of soil samples 41 3.2.3 Chemical extraction analysis of Cr speciation 44 3.2.4 SEM-EDXS 48 3.2.5 XANES experiments 48 3.2.6 FDMNES 48 3.2.7 LCF analysis 49 3.3 Results 49 3.3.1 Soil properties 49 3.3.2 Chemical extraction analysis 55 3.3.3 XRF and XANES spectra 57 3.3.4 FDMNES results and LCF analysis 59 3.3.5 SEM-EDS 65 3.4 Discussion 66 3.4.1 Soils impacted by tannery waste contamination 66 3.4.2 Cr distribution and oxidation states in soils 67 3.4.3 Cr availability in soils 69 3.4.4 Cr speciation in soils 70 vi 3.4.5 Genesis of immobile Cr(VI) in long-term contaminated soils 3.5 Summary 72 75 Chapter 4 AGING SHAPES CR(VI) SPECIATION IN SOILS: A 240-DAY INCUBATION EXPERIMENT 77 Abstract 77 Graphical Abstract 78 4.1 Introduction 79 4.2 Materials and methods 80 4.2.1 Soil sampling 80 4.2.2 Characterization of soil samples 81 4.2.3 Soil incubation 83 4.2.4 Soil Cr analysis 85 4.2.5 Kinetics modeling 87 4.2.6 FTIR spectroscopy 89 4.2.7 Soil respiration measurement 89 4.3 Results 90 4.3.1 Soil properties 90 4.3.2 Aging of Cr fractions in soils 95 4.3.3 Kinetics of available Cr(VI) in soils during aging 99 4.3.4 Changes of immobile Cr(VI) and Cr(III) fractions in soils during aging 105 4.3.5 FTIR results 106 4.4 Discussion 107 4.4.1 Transformations of Cr(VI) in soils with aging 107 4.4.2 Dominant impact factors influencing Cr(VI) transformations with aging 110 4.4.3 Multi-reaction modeling 111 4.5 Summary 119 Chapter 5 METAGENOMICS REVEALS MICROBIAL COMMUNITY IN CRCONTAMIANTED SOILS AND ITS DIRECT USE FOR MICROBAIL CR(VI) REDUCTION 120 Abstract 120 vii Graphical Abstract 121 5.1 Introduction 122 5.2 Materials and Methods 124 5.2.1 Soil sampling 124 5.2.3 Soil respiration rate 125 5.2.4 Shotgun metagenomic sequencing 125 5.2.2 Microbial Cr(VI) reduction experiment 126 5.3 Results 128 5.3.1 Active microorganisms in soils 128 5.3.2 Shotgun metagenomic sequencing results 129 5.3.3 Cr(VI) reduction by soil microorganisms 132 5.3.4 Environmental factors on microbial Cr(VI) reduction 134 5.4 Discussion 138 5.4.1 Composition of microbial communities in Cr-contaminated soils 138 5.4.2 Cr(VI) reduction by soil microorganisms 141 5.5 Summary 142 Chapter 6 CONCLUSIONS AND RECOMMENDATIONS 144 6.1 Synthesis and conclusion 144 6.2 Limitation and future research 147 Bibliography 148 APPENDIX A Density functional theory calculations 180 APPENDIX B Sequencing processing procedure used by Metagenome Bio Inc. (Waterloo, Canada). 181 APPENDIX C Sequencing data analysis by Sangon Biotech Co., Ltd. (Shanghai, China). 183 viii LIST OF TABLES Table 2.1 World reserves (shipping grade) and production (ore and concentration) of chromium by principal countries. ................................................................................................................... 10 Table 2.2 Selected chromium contamination sites reported around the globe from 2016 to 2021. ....................................................................................................................................................... 13 Table 2.3 Chemical reductants for remediation of Cr(VI) contaminated soils in the literature from 2016 to 2021. ........................................................................................................................ 21 Table 2.4 Chromate reduction rates in different organisms. ........................................................ 28 Table 3.1 Selected physical and chemical properties of contaminated soils and tannery waste used in this study. .......................................................................................................................... 50 Table 3.2 Mean of elemental contents in soils collected around three leather tanning sites in Shuitou (southern China). ............................................................................................................. 53 Table 3.3 Speciation and fractionation analysis of Cr in soils by chemical extractions. Means with standard deviations ............................................................................................................... 56 Table 3.4 Linear combination fitting (LCF) result for Cr-soil 2-1 sample system ...................... 64 Table 3.5 The weight (%) of Cr, Mn and Fe in contaminated soil sample 3-1 determined by energy dispersive spectroscopy (EDS) system ............................................................................. 66 Table 4.1 Summary of spiking treatments………..……………………………………………. 84 Table 4.2 Kinetic models used to study available Cr(VI) data in soils with aging ...................... 88 Table 4.3 Selected properties of five soils used in this study. ..................................................... 92 Table 4.4 Concentrations of available Cr, available Cr(VI) and available Cr(III) in selected samples . ........................................................................................................................................ 97 Table 4.5 Mass balance analysis of immobile Cr concentrations in mg kg-1 (standard deviation) in selected soil samples ................................................................................................................. 98 Table 4.6 The fitted coefficient (R2) and Akaike’s information criterion value (AICc) and parameters of kinetic models on available Cr(VI) aging process. .............................................. 102 Table 4.7 Fitted linear regressions of first-order and second-order reactions for Luvisol-1 and Luvisol-2 with multiple spiking of 150 mg kg-1 Cr(VI). ............................................................ 104 Table 4.8 Mean (standard deviation) of soil respiration rates in the soils. ................................ 109 Table 4.9 Isotherm constants, correlation coefficients (R2) and Akaike’s information criterion (AICc) obtained for models for Cr(VI) sorption onto soils. ........................................................ 114 Table 5.1 Selected properties (means with standard deviation, n = 3) of soil S1-6 and S3-2. .. 124 Table 5.2 A summary description of the treatments applied to K2Cr2O7 solution to determine the role of soil microorganisms in the reduction of Cr(VI) to Cr(III). ............................................. 127 ix Table 5.3 Mean (standard deviation) of soil respiration rates during 4-day incubation ............ 128 Table 5.4 Concentrations of Cr(VI) (mg L-1) and pH in 20 mg Cr(VI) L-1 of K2Cr2O7 solution after 7-day incubation at 30 oC in the presence of soil microbial consortia amended of a variety of carbon sources at 0.2 g L-1 under anaerobic condition. .......................................................... 136 x LIST OF FIGURES Figure 2.1 Eh-pH diagram for chromium .................................................................................... 15 Figure 2.2 Three-step concept of heavy metal bioavailability in soils for plants ........................ 17 Figure 2.3 The proposed Cr(VI) resistance mechanisms. ............................................................ 24 Figure 2.4 Mechanisms of Cr(VI) bioreduction under aerobic and anerobic conditions ............ 27 Figure 3.1 Particle size analysis apparatus using the pipette method. ......................................... 43 Figure 3.2 Sequential extraction results for Cr(VI) fractions (mg kg-1) in soil 1-4, soil 2-1, soil 31 and certified reference soil sample CRM041............................................................................. 47 Figure 3.3 CaCO3 content in soils (S1-3, S2-1 and S1-4) and tannery waste as determined by thermogravimetric analysis (TGA) ............................................................................................... 51 Figure 3.4 X-ray diffraction (XRD) powder patterns of the studied soil samples (S1-4, S2-1 and S3-1) and tannery waste. ............................................................................................................... 54 Figure 3.5 Detailed map of Fe, Mn and Cr distribution for soil 1-4, soil 2-1 and soil 3-1 obtained from two-dimensional (2D) mapping by using a 32-element Ge detector (Horizontal: mm; Vertical: mm) ................................................................................................................................ 57 Figure 3.6 Experimental XANES spectra of (a) Cr(VI) and Cr(III) standards; (b) Cr hot spots of soil 1-4, soil 2-1 and soil 3-1; (c) points of interest along with Cr concentration gradient in soil 21 before (1-2) and after PBE (1-4). ............................................................................................... 58 Figure 3.7 Localized distribution of Fe, Mn and Cr in soil 2-1 before and after phosphate buffer extraction (PBE)............................................................................................................................ 59 Figure 3.8 Normalized Cr K-edge XANES spectra of (a) Experimental XANES data of soil 2-1 samples and representative references; (b) FDMNES calculated spectra of CrFeO3 particle with increasing cluster sizes, 2.0 Å, 3.25 Å, 3.5 Å, 4.0 Å, 4.8 Å, 5.0 Å and 6.0 Å; (c) The modeled XANES of CrFeO3 (5.0 Å in radius) compared to the experimental data (S2-1-PBE-3). ............ 60 Figure 3.9 Cr K-edge K2-weighted EXAFS spectra: (a) representative soil 2-1 samples and FDMNES calculated spectra of CrFeO3(R3) particle with cluster sizes at 4.0 Å and 5.0 Å; (b) representative soil 2-1 samples and FDMNES calculated spectra of CrOOH (Pbmn) particle with increasing cluster sizes, 4.0 Å, 5.0 Å and 6.0 Å; (c) representative soil 2-1 samples and Cr(VI) standards; (4) Experimental low K EXAFS data of representative soil 2-1 samples and FDMNES calculated spectra of CrOOH (R3m) particle with increasing cluster sizes, 4.0 Å, 5.0 Å and 6.0 Å. ....................................................................................................................................................... 62 Figure 3.10 (a) Normalized Cr K-edge XANES spectra of modeled Cr2O3 compared to the experimental data (S2-1-PBE-3 as an example) and (b) the pre-edge. ......................................... 62 Figure 3.11 Linear combination fitting (LCF) of XANES Cr K-edge normalized spectra of soil 2-1 system before and after PBE extraction ................................................................................. 64 xi Figure 3.12 (a) Scanning Electron Microscope (SEM) images of soil 3-1; regression and correlation analysis of weight percentage of (b) Cr & Fe, (c) Cr & Mn and (d) Cr & Ca........... 65 Figure 3.13 Crystal structures and electron density maps of (a) CrFeO3 (R3) and (b) CrOOH (Pbnm)........................................................................................................................................... 71 Figure 3.14 Component weight correlation among CrFeO3, CrOOH and CaCrO4 from LCF result .............................................................................................................................................. 73 Figure 3.15 Deformation charge density of (a) CrFeO3 and (b) CrOOH with δ-MnO2. ............. 74 Figure 3.16 Simplified schematic of immobile Cr(VI) species formation in soils impacted by tannery sludge. .............................................................................................................................. 76 Figure 4.1 Incubation of soil samples in the growth chamber. .................................................... 85 Figure 4.2 CaCO3 content in soils as determined by thermogravimetric analysis (TGA). .......... 93 Figure 4.3 XRD powder patterns of the studied samples. ........................................................... 94 Figure 4.4 Changes of Cr fractions as a proportion of total added Cr(VI) during aging in Brunisol spiked with 150, 112.5 and 75 mg kg-1 Cr(VI) [(a) – (c)], in Anthrosol-1 spiked with 75, 30 and 15 mg kg-1 Cr(VI) [(d) – (f)], and in Anthrosol-2 spiked with 112.5, 75 and 30 mg kg-1 [(g) – (i)]. ............................................................................................................................................. 96 Figure 4.5 Changes of available Cr(VI) content (mg kg-1) during the whole aging period in (a) Brunisol spiked with 150, 112.5 and 75 mg kg-1 Cr(VI); (b) Anthrosol-1 spiked with 75, 30 and 15 mg kg-1 Cr(VI); (c) Anthrosol-2 spiked with 112.5, 75 and 30 mg kg-1 Cr(VI). ................... 100 Figure 4.6 Cr(VI) with aging and the fit of data by kinetic models in Brunisol spiked with 150, 112.5 and 75 mg kg-1 Cr(VI) [(a) – (c)], in Anthrosol-1 spiked with 75, 30 and 15 mg kg-1 Cr(VI) [(d) – (f)], and in Anthrosol-2 spiked with 112.5, 75 and 30 mg kg-1 [(g) – (i)]. ....................... 101 Figure 4.7 (a) Changes of available Cr(VI) content (mg kg-1) and (b) first-order plots of ln[available Cr(VI)] in Luvisol-1 and Luvisol-2 with multiple Cr(VI) spiking as a function of time. ............................................................................................................................................ 104 Figure 4.8 Changes of immobile Cr(VI) and immobile Cr(III) contents (mg kg-1) during the whole aging period in (a-c) Brunisol spiked with 150, 112.5 and 75 mg kg-1 Cr(VI); (d-f) Anthrosol-1 spiked with 75, 30 and 15 mg kg-1 Cr(VI); (g-i) Anthrosol-2 spiked with 112.5, 75 and 30 mg kg-1 Cr(VI)................................................................................................................. 105 Figure 4.9 FTIR spectra of Anthrosol-2 spiked with 112.5 mg kg-1, Luvisol-1 spiked with 450 mg kg-1 and Luvisol-2 spiked with 1500 mg kg-1 after aging. .................................................... 106 Figure 4.10 Diagram of the multi-reaction model on Cr(VI) transformations in soils. ............. 112 Figure 4.11 Experimental data and sorption isotherms of Cr(VI) onto (a) Brunisol, (b) Anthrosol-1, and (c) Anthrosol-2 (2.00 g air-dried soils in 20 mL K2Cr2O7 solution in the range of 2 – 100 mg Cr(VI)/L ............................................................................................................... 113 Figure 4.12 Simulink procedure scheme for the presented model.. ........................................... 116 xii Figure 4.13 The change of Cr concentrations (mg kg-1) in various fractions [Cr6+(l), Cr6+(ad), Cr6+(av), Cr6+(im) and Cr3+(im)] [(a) - (c)], reaction flow rates (mg kg-1d-1) [(d) - (f)] and rate coefficients of first-order kinetic model [(g) - (f)] with aging in Anthrosol-1 spiked with 75 Cr(VI) mg kg-1 and Anthrosol-2 spiked with 112.5 Cr(VI) mg kg-1. .................................................... 118 Figure 5.1 Venn diagram of microbial communities at (a) phylum, (b) class, (c) genus and (d) species levels in S1-6 in blue and S3-2 in yellow………………………………………………130 Figure 5.2 Relative abundances (%) of (b) all phyla, (d) dominant classes, (c) dominant genera and (d) dominant species in S1-6 and S3-2, determined by shotgun metagenomic sequencing. 131 Figure 5.3 Cr(VI) concentrations in K2Cr2O7 solution amended with no soil (in black), sterilized (in red) or non-sterilized Cr-contaminated soil (in blue) after 7-day incubation with or without Na-acetate under aerobic (T7 – T12) and anaerobic (T1 – T6) conditions at 30oC using (a) low Cr-contaminated soils and (b) high Cr-contaminated soils......................................................... 133 Figure 5.4 Cr(VI) removal (%) by (a)low Cr-contaminated soils (S1-2 and S1-6) and (b)high Crcontaminated soils (S2-2 and S3-2) under varying initial Cr(VI) concentration at 10 – 80 mg L-1 anaerobically (blue bar) and aerobically (yellow bar).. .............................................................. 135 Figure 5.5 Anaerobic Cr(VI) reduction by (a) S2-2 and (b) S3-2 soil microbial consortia at 30 oC at 20 – 60 mg Cr(VI) L-1 with the supplement of 0.2 g L-1 Na-acetate (black line) and at 40 mg Cr(VI) L-1 with 0.5 g L-1 Na-acetate (Red line) at incubation days. ........................................... 137 Figure 5.6 Relative abundances (%) of (a) all phyla and (b) all classes in soil samples determined by 16S rRNA sequencing......................................................................................... 140 xiii LIST OF ABBREVIATIONS AICc Akaike’s information criterion CCME Canadian Council of Ministers of the Environment CEC Cation exchange capacity CEQSS China Environmental Quality Standards for Soils CLS Canadian Light Source CSQG Canadian Soil Quality Guidelines COPR Chromite ore processing residue CTSS Corrected total sum of squares Cr(VI) Hexavalent chromium Cr(III) Trivalent chromium DFT Density functional theory calculations DWE Distilled water extraction EDS Energy dispersive x-ray spectroscopy EXAFS Extended X-ray Absorption Fine Structure FDMNES Finite difference modelling of the near-edge structure FTIR Fourier-transform infrared spectroscopy ICP-OES Inductively coupled plasma optical emission spectrometry LCF Linear combination fitting NAD(P)H nicotinamide adenine dinucleotide phosphate (NADP) + hydrogen (H); Nicotinamide adenine dinucleotide (NAD) + hydrogen (H) xiv NALS Northern Analytical Lab Services nZVI nano zero valent iron ORFs Open reading frames OTU Operational taxonomic units PBE Phosphate buffer extraction PCR Polymerase chain reaction RSS Residual sum of squares SEM Scanning Electron Microscope SPLP Synthetic Precipitation Leaching Procedure SR Soil respiration rate SIC Soil inorganic carbon SIR Substrate induced respiration SOC Soil organic mater SSR Specific soil respiration TGA Thermogravimetric analyzer TOC Total organic carbon TP Total porosity USEPA United states Environmental Protection Agency XANES X-ray near edge structure spectroscopy XRD X-ray diffraction XRF X- ray fluorescence xv ACKNOWLEDGEMENT First of all, I would like to express my gratitude and appreciation to Dr. Joselito Arocena for guiding me to science research. And I feel deep regret for his death. Prior to his passing, Dr. Joselito Arocena received a grant from Natural Sciences and Engineering Research Council of Canada, and it was transferred to Dr. Todd Whitcombe upon Dr. Arocena’s death. Without the funding, I would not have been able to do my PhD research. Additional support from the UNBC Graduate Teaching Assistantship, Graduate Conference Travel Award and Research Project Awards were much appreciated. I want to send my special thanks to Dr. Bill McGill whose guidance, support and encouragement have been invaluable throughout this study. Without his help and advice, I would not have been able to complete this research. I would also like to thank my co-supervisors, Dr. Todd Whitcombe and Dr. Mike Rutherford, and supervisory committee member, Dr. Andrea Gorrell, for their insightful comments, guidance, and discussions. Special thanks to Dr. Ning Chen, the beamline scientist at Canadian Light Source, for sharing his expertise in my research. I also thank Dr. Wei Zhang, the visiting scholar of UNBC from Changsha University of Science & Technology (China) for his continuous encouragement. I would like to express my thanks to Dr. Hualin Chen (Wenzhou University, China) for the assistance for sample sampling in contaminated sites. Many thanks to Dr. Hossein Kazemian, Erwin Rehl and Charles Bradshaw for their support and professional advice with sample analysis at UNBC Northern Analytical Laboratory Service. I would like to thank Conan Ma working at UNBC Chemstore for his support with sample delivery and lab materials supply. I extend my thanks to John Orlowsky and Doug Thompson who assisted in experiment setup and sample preparation at Enhanced Forestry Laboratory. My gratitude also extends to the graduate students at UNBC for sharing their knowledge, skills, and thoughts (Hongyuan Shi, Dr. Yuan Tian). I would also like to extend my gratitude to all my friends in China and Canada for always being supportive, inspiring, and understanding. My sincere thanks go to all my family for their unfailing love and unwavering support. xvi Chapter 1 GENERAL INTRODUCTION Heavy metal contamination of soil and the environment is a serious worldwide issue. Within the environmental literature, the simplest definition of “heavy metal” is an element with metallic properties and an atomic number > 20. However, this definition does not really encompass the issue as heavy metal contaminants are generally viewed as those elements with some level of toxicity in one or more oxidation states and/or various compounds. The most common heavy metal contaminants in the environment are chromium (Cr), copper (Cu), lead (Pb), arsenic (As), mercury (Hg), zinc (Zn), uranium (U) and cadmium (Cd).1 Globally, more than 5 million sites covering over 20 million hectare of land have been reported to be contaminated with heavy metals.2 High concentrations of heavy metal contaminants poses threats to the sustainability of ecosystems and ultimately to human health.3,4 Chromium is a heavy metal contaminant which has demonstrated to be a carcinogen as Cr(VI).5 Anthropogenic activities responsible for elevated chromium levels in the environment include metallurgical industries (steel, alloy and nonferrous alloy production), refractory industries (cement, glass, ceramics and machinery) and chemical industries (leather tanning, plating, wood preservation and pigment production).6 In the study area (Shuitou town, Wenzhou city, China) of the current work, over 30 years of continuous leather tanning activities from the 1980s to the 2010s has contaminated the soils so badly that the life of local residents is degraded.7,8 It is important to recognize that it is not the total metal concentration in soil that indicates the risk of toxicity, but rather the speciation of the metal which regulates its biogeochemical 1 behaviour.9 This is particularly the case for chromium which has two oxidation states with contrasting toxicities and mobilities: Cr(VI) is carcinogenic and highly mobile while Cr(III) is non-toxic and relatively immobile.10 Therefore, speciation information as well as the transformation of Cr in soils are critically important for developing risk-based remediation strategies. Previous studies have examined the fate of Cr through short-term spiking experiments, which emphasized the high mobility of Cr(VI) species and the prevalence of Cr(VI) reduction to Cr(III) in soils.11–17 However, transformations of trace metals is likely to continue over time spans ranging from months to years. Aging processes, such as substitution for a matrix ion, entrainment of the ion into the solid phase and surface precipitation, may lead to more extensive transformations in soils.18–21 Very little is currently known regarding chromium accumulation in long-term Cr-contaminated soils. There is a paucity of data concerning the kinetics of chromium transformations with respect to both oxidation/reduction and solubilization/immobilization in soils and in particular over extended periods of time. Considerable efforts to control the Cr(VI) concentrations below recommended levels have been made. Physical remediation are vitrification and electrokinetic remediation while chemical remediation includes immobilization techniques and soil washing.22 Nonetheless, many of these techniques require significant energy input and large quantities of chemical reagents, which may result in soil disturbance and secondary pollution.23 Alternatively, biological techniques have the advantages of minimum disturbance to the soil, low cost and simple operation.23 Bioremediation uses living organisms to convert toxic and mobile Cr(VI) to nontoxic Cr(III) and immobilized in the soil matrix, making it environmentally benign.23 At present, many microorganisms isolated from various environment samples have been reported as being capable of reducing Cr(VI) to 2 Cr(III).6,23 The unexploited potential of the indigenous microorganisms in long-term Cr contaminated soils for bioremediation still needs to be investigated to a significant extent. The purpose of this dissertation has been to examine chromium transformations occurring over a long period of time while focusing specially on the speciation of chromium in long-term contaminated soil samples. A second goal was to identify the microbial community composition in contaminated soils while evaluating the potential use of indigenous soil microorganisms for bioremediation. The results will benefit the scientific community and others studying soil fertility, land-use planning, water quality, environmental quality, soil ecology and soil bioremediation. This dissertation has five major chapters, and the contents of each chapter are briefly described below. Chapter 2 is a review of current literature summarizing occurrence of Cr in the environment with respect to its oxidation states, mobility, and toxicity, as well as remediation strategies to decontaminate polluted sites. The literature review identifies a serious gap in knowledge on the influence of time on the chromium behavior in soils. Addressing time studies on chromium chemistry in soils is the basis for much of this dissertation. The literature review also suggests exploring the use of soil microbial consortia to achieve the best bioremediation efficiency. In chapter 3, Cr speciation in three soils collected from contaminated sites was quantified and intended to highlight the genesis of immobile Cr(VI) in such soils impacted by long-term tannery activities. Chemical extraction techniques were employed to measure the total Cr contents, oxidation states, and availability in soils. Synchrotron-based x-ray fluorescence (XRF) and X-ray near-edge structure spectroscopy (XANES), along with theoretical modelling approaches were used for molecular characterization of Cr speciation in soils. Results of these assessments are 3 discussed in terms of the genesis of immobile Cr(VI) in soil and its environmental risks, as well as remediation suggestions. In chapter 4, the influence of time on Cr transformations in soils was examined. Sequential extraction methods were applied to quantify the change of various Cr species concentrations in five Cr(VI)-spiked soils for a maximum of 240 days. The kinetics of Cr(VI) transformations in soils with aging was analyzed by fitting the resulting data to chemical kinetic equations. A multireaction model was developed in MATLAB Simulink toolbox which describes various reaction flow rates among Cr fractions as a function of aging. Results of this study were discussed in terms of the competing reactions among Cr species in soils with aging and the environmental factors that govern these reactions. Chapter 5 examines whether or not indigenous soil microorganisms have evolved to have Cr(VI) reduction capacity, as a consequence of exposure to Cr-contaminated soils generated by residues from the tanning industry. Shotgun metagenomic sequencing was used to determine the difference in microbial community composition in the soil samples (designated as low and high based on the level of total Cr) collected from a long-term tannery waste contaminated area. Microbial Cr(VI) reduction efficiency was evaluated by conducting batch experiments, as influenced by electron sources, oxygen concentration and initial Cr(VI) concentration. These results are discussed with regard to potential bioremediation using mixed bacterial consortia in remediating Cr-contaminated soils or wastewater. The concluding chapter provides a synthesis of all of the above information leading to considerations for future research. This chapter discusses several themes that emerged from the previous three chapters, including the speciation of Cr in soils, kinetics of Cr biogeochemical 4 processes, and microbial Cr(VI) reduction. The final chapter also contains reflections about designing experiments, as well as new hypotheses which may drive future research. 5 Chapter 2 LITERATURE REVIEW Abstract Chromium (Cr) has been extensively released to the environment by naturally weathering of ultramafic rocks and anthropogenic activities such as mining, metal plating, tanning and wood preservation. Chromium persists as either Cr(III) or Cr(VI) in the environment. These oxidation states result in different levels of toxicity and mobilities: Cr(III) is rather benign and immobile in soils while Cr(VI) is toxic and readily transported. Therefore, developing remediation methods requires Cr speciation assessment of the sites and efficient Cr(VI) removal through physical, chemical, or biological processes. This chapter described the occurrence of Cr in the environment and its fate, toxicity, and environmental impacts. In order to conserve the environment and human health, the chemical/biological remediation processes for reducing the toxic Cr(VI) to nontoxic Cr(VI) and their efficiency have been summarised in some detail. In particular, the resistance mechanisms of soil microorganisms to Cr(VI) stress and their Cr(VI) reduction capacity as well as influential factors are discussed. 6 2.1 Introduction Environmental contamination by Cr has gained substantial consideration because of its potential toxicity and high levels in soil originating from numerous geogenic and anthropogenic activities.6 Chromium in soils can leach into groundwater and accumulate in crops grown on contaminated soils, posing health risks to humans and the environment as a result of water and food chain contamination.24 Chromium is present in two oxidation states in the environment, Cr(III) and Cr(VI), which display quite different chemical properties and affect organisms in unique ways. Cr(VI) forms anionic oxygenated ions (HCrO4- and CrO42-) which are highly soluble in soils.10 Cr(VI) is considered a group A human carcinogen because of its mutagenic and carcinogenic properties.25 In contrast, Cr(III) is an essential nutrient for humans and relatively immobile in soils.26 Therefore, knowledge of Cr speciation and transformations in soils are required to develop remediation strategies for Cr-contaminated sites. The main strategy for remediation of Cr-contaminated sites is to reduce the highly toxic Cr(VI) to less toxic Cr(III). Conventional physical and chemical methods for remediation of Crcontaminated sites include reduction, precipitation, ion exchange, and adsorption that are economically expensive and have disadvantages of high chemical reagent consumption and energy requirements.27 Microbial methods offer potential alternatives to existing technologies for Cr(VI) removal which are considered cost-effective and eco-friendly.6 Although high levels of Cr(VI) are toxic to microorganisms, many resistant bacteria and fungi species have been identified capable of reducing Cr(VI) to Cr(III) which could ultimately be employed in remediation strategies.6 7 This chapter introduces the occurrence of Cr in the environment and its fate, toxicity, and environmental impacts, especially the toxic form of Cr(VI). Some of the important efforts made to use microorganisms for potential bioremediation of Cr(VI)-polluted soils or wastewater are summarized. 2.2 Occurrence of chromium 2.2.1 Naturally geogenic Cr Naturally occurring Cr in the environment is associated with the weathering of ultramafic rocks.28,29 Chromium concentration in soils normally ranges from 20 – 200 mg kg-1 globally.30 But in serpentine soils derived from the weathering of ultramafic rocks, it typically exceeds 200 mg kg-1,31 and can reach up to 6 wt% in some ores obtained from New Caledonia.32 Serpentine soils are typically characterized as slightly acidic, contain Fe oxides (magnetite and hematite), phyllosilicates (serpentine, chlorite, and clays such as smectites and vermiculites), with a concentration of Cr (> 200 mg kg-1), Ni (> 1000 mg kg-1), and Mn (> 200 mg kg-1) exceeding the values of non-serpentine soils.31 In the Earth’s crust, ultramafic rocks and serpentinites constitute ~ 1% of the terrestrial landscape, primarily within populated areas of the Circum-Pacific and Mediterranean regions.33 Although a large number of Cr minerals are insoluble, weathering of ultramafic rocks or serpentine soils have been linked to the occurrence of elevated concentration of Cr in the surface rivers and groundwater.33–35 In a study of ultramafic rocks in the western United States, for example, concentrations up to 60 μg L-1 of total dissolved Cr was reported in observation wells in Mojave Desert in southern California,36 and decreased to 36 μg L-1 going north to Santa Cruz county.37 In a study of ophiolite-related ultramafic rocks in La Spezia (Italy), Cr(VI) 8 concentration ranged from 5 – 73 μg L-1 in 30 of the 58 groundwater samples.38 Agricultural soils derived from weathered peridotites in New Caledonia reached 700 μg L-1 of total Cr in the soil solution.39 Similarly elevated Cr has been reported in seepage soils of the Tulameen ultramafic complex of southern British Columbia.28 These groundwater concentrations can potentially leach into agricultural land leading to high levels in crops and pose a risk to human and animal population where groundwater is used as a drinking supply. 2.2.2 Anthropogenic Cr Chromium bearing ores are found in many forms, but among them only chromite in its spinel form (up to 54 wt% Cr2O3) has commercial significance.40 Total mine production worldwide was 44,000 thousand metric tons in 2019 and 40,000 thousand metric tons in 2020 (gross weight of marketable chromite core) (Table 2.1). Total world reserves are estimated at 570 million metric tons of shipping grade chromite ore with major resources located in Kazakhstan (230), South Africa (200), India (100), Turkey (26), Finland (13) and United States (0.62). 9 Table 2.1 World reserves (shipping grade) and production (ore and concentration) of chromium by principal countries. Production (unit: 1000 metric tons) Country Reservesƞ 2016 2017 2018 2019 2020 620 – – – – – Finland 13,000 – – 2,210 2,200 2,400 India 100,000 3,200 3,200 4,300 4,100 4,000 Kazakhstan 230,000 5,380 5,400 6,690 6,700 6,700 South Africa 200,000 14,700 15,000 17,600 17,000 16,000 Turkey 26,000 2,800 2,800 8,000 10,000 6,300 NA 4,160 4,200 4,250 4000 4,800 570,000 30,200 31,000 43,100 44,000 40,000 United States Other countries World total (rounded) Source: World Mineral Production, 2016-2021, USGS minerals information. Mineral Commodity Summaries.41 ƞ Reserves are defined as the supplies of the economically extractable mineral commodity. Reserves units are thousands of tons of shipping-grade chromite ore, which is deposit quantity and grade normalized to 45% Cr2O3, except for United states where grade is normalized to 7% Cr2O3 and Finland where grade is normalized to 26% Cr2O3. NA = Not available. 10 Anthropogenic Cr in soils often results from the processing of chromite ore to produce chromium compounds. Chromite ore processing residue (COPR) is an industrial waste material generated by manufacturing chromate chemicals from chromite ore. In Hudson County, New Jersey (USA), 70 years of chromate chemical manufacturing have resulted in over 130 COPR contaminated sites with reported Cr concentrations in excess of 10,000 mg kg-1.42 In southeast Glasgow, Scotland (UK), although chromate chemical works have not been in operation since 1968, Cr(VI) in the COPR-contaminated sites was still leaching out after 30 years. Groundwater in the surrounding area has been contaminated with up to 91 mg L-1.43 In other countries, such as China and India, the manufacturing of chromate has ceased due to the introduction of strict environmental regulations in recent years and new COPR contaminated sites have been reported in many places in these countries.44–47 Of the total chromite ore production, 90% is used in metallurgical industries for the production of steel, ferrous alloy, and nonferrous alloys. The remaining 10% is used in refractory industries (cement, glass, ceramics, and machinery) and chemical industries (leather tanning, plating, wood preservation and pigment).6 Leather industries are perceived to be both a consumer of Cr chemicals and a producer of pollutants. Processing one metric ton of raw hide generates 200 kg of final leather product containing 3 kg Cr, and 50,000 kg of wastewater containing 5 kg Cr.48 In 2001, Avudainayagam et al. reported that tannery waste has polluted surface soils with 62,000 mg kg-1 of exchangeable Cr in Mount Barker (Australia), as a result of leather tanning.49 In Tamil Nadu (India), the leather tannery factories are spread out across the state, resulting in high concentrations of Cr in the groundwater throughout the entire area.50,51 Bini et al. reported that soils around a leather tannery district in the Veneto region of Northern Italy contained up to 11 10,000 mg kg-1 of total Cr.52 In all of these sites, the levels of chromium contamination exceeds health limits. 2.2.3 Regulated levels worldwide Due to toxicity concerns, guidelines and regulations for Cr concentrations have been established by governments around the world for the regulation and monitoring of water and soil quality. The maximum allowable limit for total Cr in the drinking water is set at 50 μg L-1 by most of the monitoring agencies, including the World Health Organization,53 European Commission,54 Canadian Council of Ministers of the Environment,55 and Ministry of Ecology and Environment of the People’s Republic of China.56 The threshold limit for total Cr in drinking water was established at 100 μg L-1 by the United States Environmental Protection Agency.57 The acceptable level of total Cr in agricultural soils varies from country to country with Poland at 150 mg kg-1, the Czech Republic from 100 to 200 mg kg-1, Austria at 100 mg kg-1, Canada at 64 mg kg-1 and Serbia at 100 mg kg-1.24 The Canadian Council of Ministers of the Environment have recommended a 0.4 mg kg−1 limit for the toxic Cr(VI) species for agricultural soils.58 2.2.4 Reported contaminated sites worldwide Chromium contaminated soils have been reported around the globe (including urban areas such as Glasgow, Scotland, UK43 and Hudson Country, New Jersey, USA42) (Table 2.2). It has been estimated that 11% of the sites on United States National Priority List for the treatment of contaminated soils, and in Japan 14% of contaminated sites are contaminated by hexavalent chromium.59 A recent report published by Ministry of Ecology and Environment of the People’s Republic of China stated that 1.1% of soil in China had Cr contamination exceeding the risk value of 300 mg kg-1.60 12 Table 2.2 Selected chromium contamination sites reported around the globe from 2016 to 2021. Total Cr level in soil Studied sites Sources of chromium Reference 13,900 México in the city of Tultitlán, Mexico Slags and sludges from industries 61 420 – 39,000 Hunan, China Wastes from the ferroalloy industry 62 26,700 – 41,000 Hunan, China Slags from the chromate industry 63 11,200 Fuzhou, China Industrial waste 64 Tannery effluent and sludge 65 chromite ore processing residue 66 (mg kg-1) 2000 18,200 – 45,800 Edayar, Kerala, south India Kanpur, Abhaypur, Chhiwali, Godhrauli, Khanchandpur, and Rania in India 217 – 856 Schimatari area in Asopos River basin, Greece Weathering of ultramafic 3012 Odisha, India Chromite mines 398 – 2304 three areas of the Cecina Valley, Italy 67 rocks 13 68 Weathering of ultramafic rocks and serpentinites of ophiolite complexes. 69 2.3 The fate of chromium in soils 2.3.1 Speciation The concentration of total Cr in contaminated sites often exceeds the limit established by monitoring agencies. However, biological effects are not related to the total concentration of a contaminant in soils, but only to the species that are biologically available.9,70 Therefore, metal speciation information is crucial to assess risk and for remediation planning of metal polluted areas is undertaken. As will be discussed in this section, chromium speciation in soils can be quite complex as thermodynamic models the possible species to expect in a system, and speciation is also controlled by kinetics of soil reactions (e.g., sorption/desorption, reduction/oxidation, precipitation/dissolution). 2.3.1.1 Oxidation states Chromium has a broad range of oxidation states from -II to VI, but only the III and VI states are present under most conditions found in the surface environment. Figure 2.1 presents the oxidation state and hydrolytic speciation of Cr over a range of pH and Eh values. The trivalent state of Cr is present as cation species at low pH < 4 with the first and second hydrolysis products occurring from pH 4 – 6.2 and precipitation as Cr(OH)3 occurring at pH > 6.2. The hexavalent state of Cr is stable under oxidizing conditions (high Eh). All forms of Cr(VI) species are oxyanions, present as CrO42- at pH > 6.4, and as HCrO4- at pH < 6.4. 14 Figure 2.1 Eh-pH diagram for chromium from ©Dhal et al., (2013) Journal of hazardous materials, 250, page 275.6 By permission from the publisher. The oxidation state of Cr(VI) has a very high positive redox potential (e.g., > +1.0 V at pH 3) denoting that it is strongly oxidizing (Figure 2.1). Thus, chemical Cr(VI) reduction is prevalent in soils that contain electron donors.71 The reduction of Cr(VI) by soil organic matter (e.g. humic acid, fulvic acid, and humin) in both dissolved and undissolved fractions has been well documented.12,72–74 The phenolic functional group in organic matter was demonstrated to play a crucial role for scavenging Cr(VI).74,75 Other naturally occurring reductants for Cr(VI) reduction to Cr(III) include aqueous Fe(II),76,77 Fe(II)-bearing minerals,78–80 and iron sulfide minerals.81 Microorganism-induced reduction of Cr(VI) to Cr(III) takes place in three different processes: at aerobic condition, Cr(VI) reduction is associated with soluble chromate reductases that use NAD(P)H as cofactors; at anaerobic condition, microbes can use Cr(VI) as an electron accepter in the electron transport chain; reduction of Cr(VI) also take place by chemical reactions from microbial metabolism such as amino acids, vitamins, organic acids and glutathione.6 15 Under most conditions, Mn(III,IV) oxides are the only oxidants in soils capable oxidizing Cr(III) to Cr(VI) in soils since the range of pe for Mn redox transformation (12.8 - 16.7) is higher than Cr(III)/Cr(VI) transformation (10.9).82 A variety of Mn oxides have been shown to oxidize Cr(III) compounds, including pyrolusite (β-MnO2), biseriate, δ-MnO2, birnessite, and Mn biooxides from bacterial Mn(II) oxidation.83–88 The rate of Cr(III) oxidation by Mn(III,IV) oxides is often proportional to the Cr(III) solubility from different Cr(III) sources such as Cr(III) citrate complexes, Cr(OH)3 and CrxFe1-x(OH)3.89–91 2.3.1.2 Environmental availability The term speciation is a multi-faceted term.9 Metal speciation also refers to the chemical form of the metal in the soil solution, either as a free ion or complexed to a ligand, in the gaseous phase, and distributed amongst solid phases within the soil.9 Organisms respond only to the bioavailable fraction of a contaminant in soils.70,8 Bioavailability was described in three conceptual steps (Figure 2.2)92: (i) environmental availability, which is determined by an available amount of the total content in the soil including the actual available fraction dissolved in the pore water and the potential available fraction adsorbed on the soil matrix; (ii) environmental bioavailability, which is defined as the fraction of dissolved metal species in the pore water which can be taken up by plants; and (iii) toxicological bioavailability, which refers to a physiological induced bioaccumulation or biological effect of heavy metals within organisms which depends on a number of complex processes, such as translocation, metabolism, and detoxification. Throughout this dissertation, environmental availability was used to indicate the risks of Cr contamination in soils. 16 Figure 2.2 Three-step concept of heavy metal bioavailability in soils for plants from ©Kim et al. (2015) Environmental geochemistry and health, 37(6), page 1045.92 By permission from the publisher. The two oxidation states of Cr exhibit the opposite availability in soils.10 Cationic Cr(III) complexes are relatively immobile due to sorption on negatively-charged soil colloids and the formation of sparingly soluble complexes with hydroxide, organic ligands and polymers.93 Soils with higher quantities of clay, inorganic carbon, higher pH and higher cation exchange capacity generally sequester more Cr(III).94 Anionic Cr(VI) is highly available and can only be adsorbed on positively charged Fe(III) (hydr)oxides at low soil pH.10 Cr(VI) adsorption increases with increasing total soil organic carbon and decreasing soil pH.71 17 2.3.2 Aging processes Soil biogeochemical reactions occur over a wide time scale, ranging from microseconds and millisecond for ion association, ion exchange, and sorption reactions to years for mineral solution (precipitation/dissolution reactions) and mineral crystallization reactions.95 The adsorption of chromate on goethite over the pH range of 6.5 – 7.5 occurs on millisecond time scales.96 The reduction of Cr(VI) to Cr(III) by aqueous Fe(II) at 25oC from pH 2.0 – 10.0 is completed within 5 minutes.97 The rate of Cr(VI) reduction by humic acids has been shown to be on the order of days at pH 2 – 5. Aging processes include substitution for a matrix ion (i.e., lattice penetration), entrainment of the ion into the solid phase (i.e., recrystallization), surface precipitation, and diffusion into micropores.98 These processes may lead to prolonged transformations of metals in soils. For example, Liang et al. (2014) reported that spiked arsenic in soils was transformed from an available form and incorporated into amorphous and crystallized Fe/Al bound As after 33 weeks.20 Wang et al. (2017) reported that after a one-year incubation period, soluble selenium accumulated in the residual fraction in acidic soils.21 2.4 Environmental impacts 2.4.1 Toxicity of Cr(VI) and Cr(III) According to the World Health Organization (WHO), Cr(III) is not considered to be carcinogenic, but Cr(VI) has been known to be a carcinogen for more than a century.99 The US Environmental Protection Agency (EPA) has classified Cr(VI) as a Group A human carcinogen since 1986.100 Cr(VI) toxicity is related to its easy transport across cell membranes in both prokaryotic and eukaryotic organisms followed by subsequent Cr(VI) reduction in cells.101 Cr(VI) uptake by organisms is an active process via a phosphate or a sulfate transporter because of the structural resemblance of CrO42- with phosphate and sulfate ions.101 After Cr(VI) enters into cells, 18 short-lived intermediates (e.g., reactive oxygen species) released from Cr(VI) reduction induce carcinogenesis such as increased errors during RNA transcription, induction of oxidative stress and DNA oxidation.102,103 Massive exposure to Cr(VI) has been firmly associated with lung cancer.104 In contrast, Cr(III) is an essential nutrient for humans that helps with glucose and lipid metabolism.26 2.4.2 Toxicity to the environment High concentrations of chromium in soil can lead to significant losses in soil fertility and decreases in agricultural yields,105 as well as disrupting the activity of indigenous soil microorganisms.106 Pradhan et al. (2019) reported that organic carbon, available nitrogen and available phosphorus were significantly lower in Cr-contaminated soil as compared to control soil.107 Sun et al. (2019) reported that high concentrations of Cr(VI) inhibited the growth of nitrification and denitrification microbial communities and reduced the overall diversity of the bacterial communities.108 Oruko Ongon’g et al. (2020) reported that seed germination (in percentage of total planted seeds) of five vegetables was inhibited grown on Cr(VI)contaminated soils, decreasing from 16% when spiked with 23 mg kg-1 Cr(VI) to 1% when spiked with 456 mg kg-1 Cr(VI).109 Chen et al. (2014) reported that average Cr contents in collected vegetables (bokchoy and celery) from contaminated sites (261 – 2177 mg kg-1 total Cr in soil-1) were 18.5 and 18.9 mg Cr kg-1. Since lettuce, bokchoy and celery are dietary vegetables for local residents in the study area, consumption of Cr-enriched vegetables poses high health risks.8 19 2.5 Bioremediation of Cr-contaminated sites 2.5.1 Remediation strategy by reduction Unlike organic contaminants, metals do not biodegrade or decay in soils. Reduction of toxic and mobile Cr(VI) to nontoxic and immobile forms of Cr(III) is a key strategy for the remediation of Cr(VI) containing soil to meet regulatory standards and protect human and environmental health.110 Reduction methods can be roughly divided into electrochemical, chemical and biological reduction. Electrochemical reduction refers to the technology of applying a low direct electric voltage gradient through a pair of electrodes inserted into contaminated soils to promote migration of Cr(VI) anions toward the anode for reduction.111 Electrochemical reduction is rapid, allowing it to be applied for in situ soil remediation. The use of urea-titania nanotubes as cathodes removed 97% of 100 mg L-1 Cr at 5V within 15 min.112 But its efficiency is sensitive to soil properties such as permeability and the high economic and energetic cost limit its widespread application.113 Chemical reduction has been a widely used approach because of its high efficiency and adaptability.114 Common used reductants for chemical remediation of Cr(VI) contaminated soils include iron-based reductants, sulfur-based reductants, and organic amendments (Table 2.3).114 Adding FeSO4 solution to Cr(VI)-spiked soil adopting a 30:1 molar ratio between Fe(II) and Cr(VI) completely removed 96 mg kg-1 of total Cr(VI) after 40 hours of the treatment.115 A dosage of 5% Na2S was applied to Cr(VI) contaminated soil by a mechanochemical treatment, decreasing the total leached Cr(VI) from 664 to 0.84 mg L-1 in 6 hours.116 In spite of the high Cr(VI) removal efficiency, the cost of synthesis for chemical reductants is too high to be used on large-scale remediation,27 and most of them are only economically viable at high or moderate concentrations of metals but not at low concentrations (1 – 100 mg L-1).117 The use of chemical 20 reductants may cause secondary pollution or other environmental risks.27 For example, the amendment with nano zero valent iron (nZVI) reduces microbial biomass and activities in soils in which microbial populations were already stressed by the contamination.118 Table 2.3 Chemical reductants for remediation of Cr(VI) contaminated soils in the literature from 2016 to 2021. Reductants Source of contamination Initial concentrations Dosage Performances Reference Iron-based reductants nZVI Cr(VI)-spiked soil Total Cr: 86 mg kg-1 2.5% (w/w) Cr concentration decreased to 7 mg kg-1 119 S-ZVIbm Cr(VI)-spiked soil Water soluble Cr(VI): 17.5 mg L-1 5% (w/w) A complete removal with 3 hours 120 5% (w/w) 99.5% of total Cr(VI) was reduced 121 FeS-ZVIbm Cr(VI)-spiked soil nZVIbiochar Cr(VI)-spiked soil Total Cr(VI): 448 mg kg-1; Water soluble Cr(VI): 440 mg L-1 Total Cr(VI): 320 mg kg-1 8 mg g Immobilization efficacy of Cr(VI) was 100% -1 Total Cr(VI): 795 mg kg-1; 8% (w/w) A complete removal of leached Cr(VI) in 120 minutes 123 -1 Immobilization efficacy of Cr(VI) was 99% 124 98.7% of Cr(VI) immobilization 125 nZVI-rice husk biochar Cr(VI)-spiked soil nZVI-VR Cr(VI)-spiked soil Total Cr(VI): 198 mg kg-1 50 mg kg nZVI/Ni Cr(VI)-spiked soil Total Cr(VI): 119 mg kg-1 50 mg kg-1 Leached Cr(VI): 62.4 mg L-1 21 122 Table 2.3 Chemical reductants for remediation of Cr(VI) contaminated soils in the literature from 2016 to 2021 (continued). Reductants Source of contamination Initial concentrations Dosage Performances nZVI/Cu Cr(VI)-spiked soil Total Cr(VI): 200 mg kg-1 12 mg kg-1 99% of Cr(VI) reduction at pH 5 126 5% (w/w) Leached Cr(VI) decreased to 0.84 mg L-1 116 127 Reference Sulfur-based materials Total Cr(VI): 13806 mg kg-1 Na2S A chromium salt factory in Shandong, China CaS5 A planting site in Beijing, China Leached Cr(VI): 115 mg L-1 3% (w/w) Leached Cr(VI) decreased to 0.51 mg L-1 0 Shuitou, China Total Cr(VI): 12.6 – 42.5 mg 4.0 mg g-1 kg-1 PBR Cr(VI) dropped to < 0.4 mg kg-1 128 *CMC-FeS Cr(VI)-spiked soil Total Cr(VI): 1407 mg kg-1 1.5:1 (molar ratio) 98% of Cr(VI) was reduced 129 CMC-nFeS An industrial site in Chongqing, China Total Cr(VI): 205 mg kg-1 45% of pore volumes CMC-nFeS Cr(VI)-spiked soil Total Cr(VI): 56 - 502 mg kg-1; CMC-FeSbiochar Cr(VI)-spiked soil Leached Cr(VI): 11.9 mg L-1 S Leached Cr(VI): 664 mg L-1 Leached Cr(VI) decreased from 45.8 to 0.05 mg L- 0.1 g L-1 2.5 mg g 22 130 1 -1 Reduction capacity was 55 – 199 mgCr(VI) g-1 FeS 131 Leached Cr(VI) reduced to 0.63 mg L-1 132 Table 2.3 Chemical reductants for remediation of Cr(VI) contaminated soils in the literature from 2016 to 2021 (continued). Reductants Source of contamination Initial concentrations Dosage Performances Reference 133 Organic amendments A municipal solid Cr(VI)-spiked organic soil waste Total Cr(VI): 300 mg kg-1 10 – 20% (w/w) Cr(VI) decreased from 65.9 to 15.3 mg kg-1 and from 96.3 to 21.1 mg kg-1. Biogas solid Cr(VI)-spiked residue soil Total Cr(VI): 300 mg kg-1 150 g kg-1 52.1% of Cr(VI) reduction 134 Chromium residue in Henan, China Total Cr(VI): 2629 mg kg-1 1:1 (w/w) Cr(VI) decreased to 62 – 146 mg kg-1 after 50 days. 135 5% (w/w) Cr(VI) reduction at 75 mg kg-1 in soil 1 and 88 mg kg-1 in soil 2 136 Decreased the leaching amount of Cr by 47% and 55% 137 Leachability of Cr was reduced by over 81% 138 Compost Biochar Cr(VI)-spiked soil Total Cr(VI): 100 mg kg-1 Manure and biochar An abandoned chemical Total Cr(VI): plant in 150 mg kg-1 Sichuan, China Biochar/Fe composite An abandoned electroplating Total Cr(VI): company in 6622 mg kg-1 Fujian, China 9 g kg-1; 7 g kg -1 5% (w/w) *CMC-Carboxymethyl cellulose-stabilized. Biological reduction refers to the use of microorganisms capable of Cr(VI) reduction in detoxifying Cr-contaminated soils. Bioreduction has the advantages of causing little disturbance to soil, convenient operation, low cost and less secondary pollution.22 Bioremediation will be discussed in the following sections. 23 2.5.2 Microbial resistance Many microorganisms have the potential to survive toxic-metal-polluted environments by developing mechanisms to avoid toxicity.139 The resistant mechanisms, including reducing uptake, extracellular Cr(VI) reduction, biosorption, reducing oxidative stress, Cr(VI) efflux, intracellular Cr(VI) reduction, bioaccumulation, and DNA repair, are depicted in Figure 2.3.140 Figure 2.3 The proposed Cr(VI) resistance mechanisms from ©Xia et al. (2021) Journal of Hazardous Materials, 401, 123685, page 4.140 By permission from the publisher. Cr(VI) uptake and efflux Cells do not have specific pathway for the uptake of Cr(VI), but Cr(VI) can easily pass-through cell membranes via sulfate transporters due to the structural resemblance of sulfate ion (SO42-) and chromate ion (CrO42-).141 However, sulfate transports in bacteria such as Caulobacter crescentus CB15 N are downregulated under Cr(VI) stress, which 24 implies the bacteria can close their SO42- transport pathway to reduce the uptake of Cr(VI).142 Efflux is another strategy for microorganisms to resist Cr(VI). For example, chromate resistance in both Cupravidus metallidurans and P. aeruginosa was via the ChrA efflux protein with produced resistance levels being 4 and 0.3 mM, respectively.143 Cr(VI) reduction by microbes Extracellular Cr(VI) reduction could be regulated by soluble proteins exported to the extracellular medium, protecting microbes from the Cr(VI) toxicity by minimizing its active intracellular transport.144 Intracellular Cr(VI) reduction was mediated by soluble reductase enzymes such as ChrR, YieF, CsrF, FerB, NemA, FrP, NfoR, etc., using NAD(P)H as the electron donor.140 More details of microbial Cr(VI) reduction were discussed in the section 2.5.3. Reducing oxidative stress caused by Cr(VI) and DNA repair system Since the generation of reactive oxygen species (ROS) occurs during the Cr(VI) reduction to Cr(III), the participation of bacterial proteins in the defense against oxidative stress induced by Cr(VI) is considered another mechanism of Cr(VI) resistance.145 Ackerley et al. (2006) reported that the growth of Escherichia coli K-12 ceased within 3 hours of chromate exposure and partially recovered by 5 hours. The increased levels of proteins such as SodB and CysK, can counter oxidative stress, suggesting a cellular defence against the Cr(VI) toxicity in Escherichia coli K-12.146 High levels of oxidative stress induced by Cr(VI) results in DNA damage, and DNA damage repair was another important mechanism for bacteria to maintain their functions.145 Caulobacter crescentus showed the up-regulation of genes related to repair of DNA damage (endonucleases, RecA protein) in response to Cr(VI).142 Another DNA repair system RuvRCAB was found in Alishewanella sp. WH16-1 that contributes to Cr(VI) resistance.147 25 Biosorption and bioaccumulation of Cr(VI) Biosorption and bioaccumulation reduce the Cr(VI) mobility and thus its toxicity. Cr(VI) was found to combine with the functional groups on the cell surface in Cyclotella sp. to detoxify and was transported into the cell by binding to the carrier protein.148 In the fungus Neurospora crassa, the uptake of Cr (VI) is by CHR-1 and accumulated in vacuolar system to enhance Cr(VI) resistance.149 2.5.3 Microbial Cr(VI) reduction A wide diversity of microorganisms is known to have evolved to be capable of reducing Cr(VI) to Cr(III). Microbial Cr(VI) reduction is found to be cometabolic (not participating in energy conservation) at aerobic condition and predominantly dissimilatory under anaerobic conditions.150 As shown in Figure 2.4, aerobic Cr(VI) reduction is generally associated with soluble reductases and requires NAD(P)H as an electron donor.140 The most well studied chromate reductase so far has been ChrR, a soluble flavin mononucleotide (FMN) binding enzyme able to catalyze the electron transfers from NAD(P)H to Cr(VI).151,152 The ChrR reductases have been identified in various bacteria, such as Pseudomonas putida151, Bacillus sp. DHS-12(7)153, Stenotrophomonas maltophilia154 and Escherichia coli155, which exhibits great performance of Cr(VI) reduction. Other identified reductases in bacteria include YieF, FerB, NfsA and NfsB, NemA, AzoR, Frp, YcnD, NfoR and CsrF.140 Under anaerobic condition, Cr(VI) reduction is usually associated with membrane bound reductases such as flavin reductases and cytochromes that can be part of the electron transport system and use Cr(VI) as the terminal electron accepter (Figure 2.4). The first case of microbial reduction of Cr(VI) was reported by Romanenko & Koren’Kov (1977).156 The strain of Pseudomonas dechromaticans, isolated from sewage sludge, reduced Cr(VI) under anaerobic condition. The Cr(VI) reductase activity was found to be associated with the cytoplasmic 26 membrane of anaerobically grown Shewanella putrefaciens MR-1, with formate and NADH served as electron donors.157 Anaerobic Cr(VI) reduction by Shewanella oneidensis MR-1 utilized H2 as electron donor and Cr(VI) as an electron acceptor in the electron transport chain.158 Figure 2.4 Mechanisms of Cr(VI) bioreduction under aerobic and anerobic conditions (MR: membrane bound chromate reductase; SR: soluble chromate reductase) from ©Tang et al. (2021) Ecotoxicology and environmental safety, 208, 111699, page 5.159 Adapted with permission from the publisher. Microbial Cr(VI) reduction to Cr(III) can occur indirectly. Iron-reducing bacteria or sulfatereducing bacteria anaerobically produce soluble Fe(II) or H2S which can react with Cr(VI) to form insoluble Cr(III) complexes.6 Smith & Gadd (2000) reported that mixed culture sulfatereducing bacterial biofilms reduced sulfate to sulfide using lactate as a carbon source and the biogenic sulphide reduced 88% of the 26 mg L-1 Cr(VI) within 24 hours.160 Bishop et al. (2014) reported that structural Fe(III) in clay minerals was bioreduced to Fe(II) by an iron-reducing 27 bacterium Geobacter sulfurreducens in the presence of acetate, and biogenic Fe(II) effectively reduced Cr(VI) in a stoichiometric ratio close to 3 at 30 oC.79 Table 2.4 lists the reported Cr(VI) resistant and reducing bacteria isolated from various polluted environments (Table 2.4). Salamanca et al. (2013) documented a Cr(VI) reducing bacterial strain Pseudomonas aeruginosa CRM100 isolated from a wastewater treatment plant that could reduce 99.8% of Cr(VI) at 100 mg L-1 in 7 days in the presence of citrate in anaerobic systems.161 Das et al. (2014) reported that the bacterial strain Bacillus amyloliquefaciens isolated from chromite mine soils exhibited high tolerance to Cr(VI) up to 900 mg L-1 with the fastest reduction rate of 2.22 mg Cr(VI) L-1 h-1 observed for 100 mg L-1 Cr(VI), pH 7 and 35oC under aerobic condition.162 Table 2.4 Chromate reduction rates in different organisms. Microorganisms Lysinibacillus fusiformis ZC1 Intrasporangium sp. Q5-1 Bacillus cereus Sources of isolation Electroplating factory wastewater Crcontaminated soil Cr(VI) contaminated soil Isolation medium Cr(VI) Con. (mg L-1) Cr(VI) reduction rate (mg L-1 h-1) Condition Reference Aerobic; Sodium acetate 52 4.16 pH 8.0; 163 37oC Aerobic; Acetate, NADH 52 1.98 pH 8.0; 164 37oC Aerobic; Glucose 60 0.83 pH 5 – 7; 37oC 28 165 Microorganisms Pseudomonas aeruginosa CRM100 Sources of isolation Waste soil samples and sludge Isolation medium Cr(VI) Con. (mg L-1) Cr(VI) reduction rate (mg L-1 h-1) Condition Reference Aerobic; Citrate 100 0.5 pH 7.5; 161 37oC Aerobic; Bacillus sp. CRB-B1 Sludge samples Fructose 100 4.17 pH 6 – 8; 166 37oC Enterobacter sp. DU17 Staphylococcus aureus K1 Rhizosphere of tannery waste Glucose dump site Metal contaminated industrial effluent Aerobic; 100 4.17 pH 7.0; 167 37oC Aerobic; Tryptic Soy Broth 100 4.17 pH 8.0; 168 35oC Aerobic; Bacillus sp. M6 Ferroalloy Company site Glycerol 200 1.28 pH 7.0; 169 37oC Stenotrophomons sp. WY601 Ochrobactrum intermedium Rb2 Ochrobactrum intermedium SDCr-5 Tannery sludge Tannery water sample Anaerobic; Lactose, fructose and glucose 300 Gluconate 1,000 4.15 pH 7.0; 170 37oC 3.80 Aerobic; 171 37oC Aerobic; Effluent of dying industry Sodium acetate 1,500 12.9 pH 8.0; 37oC 29 172 2.5.4 Factors affecting microbial Cr(VI) reduction The efficiency of microbial Cr(VI) reduction is limited to the availability of various chemical factors, including electron donors and electron mediators (or electron shuttles).173 The oxidation of electron donors enables the supply of electrons to Cr(VI) to facilitate the reduction process. The most common electron donors for Cr(VI) reduction are organic molecules (glucose, fructose, lactose, pyruvate, lactate, citrate, glycerol, acetate, formate, NADH/NADPH, reduced glutathione, etc.) in which oxygen-containing functional groups (C-O, COOH, C-OH and C-O-R) are present.173 The enzyme CsrF from Alishewanella sp.WH16-1 efficiently catalyzed Cr(VI) reduction using NADPH as its electron donor.174 Arthrobacter sp. Sphe3 could completely reduce 45 mg L-1 Cr(VI) at the condition of 30oC, pH 8 and 10 g L-1 of glucose.175 Anaerobic reduction by Pannonibacter phragmitetus LSSE-09 is promoted more in the presence of acetate and lactate rather than glucose.176 Anaerobic Cr(VI) reduction by sulfate reducing bacteria has been shown to perform better using ethanol than glucose.177 These studies suggest that the utilization of electron donors for Cr(VI) reduction is species-dependent, resulting in various reduction rates. Electron mediators accelerate the Cr(VI) reduction activity by shuttling from low-potential electron donors to mediators and then from low-potential mediators to electron acceptor which is Cr(VI).178 Various humic substances and their quinoid analogs such as lawsone, menadione, anthraquinone, anthraquinone-1-sulfonate, anthraquinone 2-sulfonate, anthraquinone-1,5disulfonate, anthraquinone-2,6-disulfonate, anthraquinone-2,7-disulfonate, 1chloroanthraquinone, 1,5-dichloroanthraquinone, 1,4,5,8-tetrachloroanthraquinone, etc., have been popularly used to promote the electron transfer rate for microbial Cr(VI) reduction.178 Guo et al. (2012) reported that with initial Cr(VI) concentration of 32 mg L-1, Cr(VI) reduction 30 efficiency by Escherichia Coli BL21 was 98.5% in the presence of anthraquinone 2-sulfonate, and decreased to 21 – 34% in the absence of anthraquinone 2-sulfonate.179 Huang et al. (2019) reported that adding anthraquione-2,6-disulfonate or humic acid as a mediator doubled the microbial Cr(VI) reduction rate. Cr(VI) reduction capability decreases with the initial concentration due to the reduced cell density and bacterial activities at higher concentration of Cr(VI).180 In a study by Ma et al. (2019), the Cr(VI) reduction rate by a mixed bacterial consortium collected from a Crcontaminated soil declined as the Cr(VI) concentration rose from 2 – 50 mg L-1, with the decreasing extracellular protein concentrations in the medium. The Cr(VI) removal in 24 hours by Bacillus sp. CRB-B1 reached to 86.2% at 200 mg L-1 Cr(VI) but decreased to 43.1% at 300 mg L-1 Cr(VI), indicating the reduction capacity was jeopardized with high Cr.181 The removal percentage of Asperillus flavus CR500, isolated from electroplating wastewater, decreased with increases in Cr(VI) concentration because of decreasing biomass.182 The pH can affect the availability of Cr(VI) ions, the activity of reductase and the binding sites of Cr(VI) on cellular surfaces, thereby affecting microbial Cr(VI) removal efficiency.183,166,184 Many Cr(VI) detoxification studies mediated by microorganisms have been reported at neutral/near-neutral pH (Table 2.4). The strain of Bacillus sp. CRB-B1 showed 100% Cr(VI) removal efficiency at 100 mg L-1 Cr(VI) within 24 hours at a pH between 6 – 8. As pH decreased to 4 or increased to 10, Cr(VI) removal was severely inhibited with Cr(VI) removal decreasing to < 40%.166 The optimized Cr(VI) reduction by Sporosarcina saromensis M52 occurred at pH 7.5 and 30 oC, and the bacterial growth was hampered under acidic conditions.185 Mangaiyarkarasi et al. (2011) reported that an alkaliphilic Bacillus subtills isolated from tannery effluent 31 contaminated soil could grow and reduce 96% of 50 mg L-1 Cr(VI) at pH 9, three times higher than that at pH 6.183 As shown in Table 2.4, the temperature of microbial reduction typically carried out in the range of 30 – 40 oC. Beyond the optimum temperature, cell growth is significantly affected and thus can be detrimental for Cr(VI) bioreduction.186 Tan et al. (2020) reported that Cr(VI) reduction was observed at all temperatures between 25 and 41oC, with 33 – 37oC being the optimal temperature for effective Cr(VI) reduction by Bacillus sp. CRB-B1.181 Cr(VI) removal efficiency dropped from 100% at 33oC to 44.9% at 25oC at 100 mg L−1 Cr(VI).181 Das et al. (2014) reported that 35oC was the optimum temperature for Cr(VI) reduction by Bacillus amyloliquefaciens. Temperatures above 40oC or below 30oC dramatically decreases bacterial growth and consequently the Cr(VI) reduction efficiency.162 Low temperatures reduce the fluidity of cell membranes and affect the function of the transport system which prevents substrates entering into cell for cell growth. High temperatures cause loss of chromium reductase function, alteration of membrane structure, and inactivation of protein synthesis mechanisms.186 Thermophilic microorganisms surviving under elevated temperatures (~65 – 70oC) were reported to reduce Cr(VI) to Cr(III), which offers remediation in high temperature environments.187 Reduction of hexavalent chromium by the thermophilic methanogen Methanothermobacter thermautotrophicus was observed at up to 70oC with complete reduction of 20 mg L-1 Cr(VI).188 Naturally contaminated sites are generally characterized by the coexistence of a large number of toxic and nontoxic cations and therefore other metals may have influence on Cr(VI) bioreduction.186 Bhattacharya and Gupta (2013) reported Cr(VI) reduction by Acinetobacter sp. B9 was only 34.6%, 10.3% and 9.0% of Cr(VI) removal in the presence of Ni, Zn and Pb, respectively.189 Significant decreases in Cr(VI) reduction efficiency (58.4 – 70.1%) was also 32 observed by Enterobacter sp. DU17 at 100 mg L-1 in the presence of 50 mg L-1 Co2+, Cd2+ and Ni2+, with encouraging increases in Cr(VI) reduction efficiency (100%) in the presence of 50 mg L-1 Cu2+.167 Increases in the reduction could be attributed to the Cu2+ acting as an electron redox center for shuttling the e- between different protein subunits.190 Owing to the co-existence of oxyanions and Cr(VI) in natural wastewater or groundwater, Cr(VI) removal efficiency in presence of Oxyanions such as PO43-, SO42-, SO32-, and NO3- have also been investigated.186 He et al. (2015) reported that Cr(VI) and nitrate were simultaneously removed by Pseudomonas aeruginosa PCN-2 under aerobic conditions, suggesting the competition between Cr(VI) and NO3- for the electron from the bacterial electron transport chain. Ma et al (2019) reported that Cr(VI) reduction by a mixed bacterial consortium was unaffected in the presence of SO42- and HCO3- , but was partially limited by the existence of NO3- at concentrations 10 and 50 mg L-1, with 1% and 7% inhibition of Cr(VI) reduction, respectively.191 2.5.5 Pure cultures or mixed bacterial consortia The bioreduction of Cr(VI) to Cr(III) by pure cultures has been commonly reported as a potential bioremediation strategy in bioreactors. However, using pure cultures for remediation is sensitive to the environmental conditions, which limits the practical application in the field.192 It has been suggested that the survival and stability of microorganisms are better when they are present as a mixed consortium, which may be more suitable for field applications.193 Molokwane et al. (2008) reported purified individual species did not achieve the same level of Cr(VI) reduction as observed in mixed cultures from sludge, indicating possible synergy of interspecies interactions necessary for optimum Cr(VI) reduction.194 Qu et al. (2018) demonstrated the alternating growth between a mixed microbial consortium of Geotrichum sp. and Bacillys sp. in bioreactor with pH changes greatly improved the bioreduction efficiency.195 Aparicio et al. (2018) 33 reported that four actinobacteria strains (Streptomyces sp. M7, MC1, A5 and Amycolatopsis tucumanensis DSM 45259) achieved the best removal of Cr(VI) and lindane in both liquid media and soil, while single species could not effectively achieve removal.196 2.5.6 Biostimulation or bioaugmentation The bioremediation within an environmental matrix (e.g. soil, sediment, sludge or wastewater) can be improved either by stimulation of the indigenous microorganisms by addition of nutrients or electron acceptors (biostimulation) or by the introduction of isolated strains with desired catalytic abilities which assist the local population (bioaugmentation).197 Successful attempts made to bioremediate the Cr-contaminated soils using native microbial species are presented here. Chai et al. (2009) reported that nutrient addition to Cr-contaminated soils result in the decrease of total Cr(VI) from 463 to 10 mg kg-1 in 10 days aerobically but did not change the total Cr(VI) concentration in autoclaved soils, indicating Cr (VI) remediation in tested soils was due to Cr(VI) reduction by indigenous microorganisms.198 Leita et al. (2011) reported that the addition of glucose increased the rate of aerobic Cr(VI) reduction in soil by promoting the growth of indigenous microbial biomass.199 In a soil bioreactor, 97% of Cr(VI) was reduced within 20 days at a bacterial a concentration of 15 mg g-1 soil-1 enriched using 50 mg molasses g1 soil-1 as a carbon source in the presence of mineral medium.200 Bioaugmentation of a membrane bioreactor with Aeromonas hydrophila LZ-MG14 removed 93.7% of 0.5 mmol L-1 Cr(VI) with 12 hours with a significant increase in biomass.201 2.5.7 The fate of Cr(III) The Cr(III) end-products from microbial Cr(VI) reduction ultimately determine the bioavailability of Cr(VI).202 During the reduction, produced Cr(III) can combine with the anions in the medium to form precipitates or complexes, thereby accumulating intracellularly or 34 extracellularly.159 For example, Cr(VI) reduction by Bacillus cereus XMCr-6 formed soluble Cr(III) complexes through interactions with small organic molecules and the Cr(III) could also bind to cells by coordination with functional groups on the bacterial surfaces.203 Serratia sp. C8, isolated from tannery sediments, can reduce Cr(VI) to Cr(III) by intracellular reductase, causing the accumulation of Cr(III) in the cell.204 A Bacillus strain TCL uses exopolysaccharide (EPS) as sites to combine with Cr(III) with the formation of extracellular complexes of EPS.205 Research has been done on improving the Cr(III) immobilization after reduction. Owing to the preferential coordination of Cr(III) to the soluble organic molecules in the bacterial culture medium, Cheng et al. (2010) proposed to separate harvested bacteria from residual soluble organic molecules by filtration, in order to achieve better immobilization of Cr(III) by bacterial cells.206 Pan et al. (2014) used planktonic cells and biofilms of Bacillus subtilis ATCC-663 for Cr(VI) reduction and Cr(III) immobilization, respectively. The results showed 100 mg L-1 Cr(VI) decreased to 0 after 72 hours and 100% of Cr(III) was immobilized in the following 72 hours.180 2.6 Summary This chapter reviewed the chromium occurrence, use, speciation, toxicity and environment impacts. Naturally weathering of ultramafic rocks and extensive industry uses of Cr (e.g., chromite ore processing residue, leather tanning wastes) lead to the increased of Cr concentration in the environment. Chromium mainly exists as two chemical forms, Cr(III) and Cr(VI). The speciation, mobility, toxicity differ significantly with its chemical speciation. Bioremediation has great promise to formulate an acceptable strategy because of microbial reduction of toxic Cr(VI) to nontoxic Cr(III) and is both environmentally friendly and economical. The mechanisms involved in the process of strain resistance and how some abiotic factors (carbon sources, electron mediators, initial Cr(VI) concentration, pH, temperature, co-existing ions) affect the rate 35 of microbial Cr(VI) reduction were reviewed. Based on the data summarized in the review article, the following research gap need to be explores: (i) a detailed research is required about the transformations of various Cr species in soils with aging; (ii) the capability of indigenous mixed soil microorganisms from Cr-contaminated environment samples in reducing Cr(VI) to Cr(III) is to be quantified and the optimal conditions are to be defined for bioremediation in Cr(VI)polluted environments. 36 Chapter 3 FORMATION AND GENESIS OF IMMOBILE CR(VI) IN LONG-TERM CONTAMINATED SOILS Abstract Chromium speciation in naturally contaminated soils appears more complex than spiked studies have shown. This study characterized Cr speciation (oxidation states; availability; molecular geometry) intended to highlight the genesis of immobile Cr(VI) species in long-term tannery waste contaminated soils. In a series of samples obtained from Shuitou in China, chemical extraction methods showed Cr(III) was dominant (> 96.7% of total Cr at 4320 – 27200 mg kg-1), with Cr(VI) concentration up to 144 mg kg-1. Of the total Cr(VI) present, immobile Cr(VI) represented > 90%. Synchrotron-based X-ray near edge structure spectroscopy (XANES) showed the occurrence of Cr(VI), which was not removed by phosphate buffer extraction, confirming significant amount of immobile Cr(VI) fractions in soils. X-ray fluorescence maps exhibited the heterogeneous distribution of Cr in soils associated with both Mn and Fe. Such a distribution suggests Cr(III) oxidation to Cr(VI) by Mn oxides and a possible immobilization of both Cr(III) and Cr(VI) on to Fe (hydr)oxides. Linear combination fitting of XANES spectra revealed fractional weights (%) in samples were CrFeO3 (49.3 – 53.6), CrOOH (22.3 – 29.5) and CaCrO4 (13.2 – 25.3). The results demonstrate i) Cr(VI) is immobilized in soils; ii) mechanisms for Cr(VI) immobilization is via precipitation as CaCrO4 and via recrystallization with Fe(hydr)oxides. This work has been published as Shi, J., McGill, W. B., Chen, N., Rutherford, P. M., Whitcombe, T. W., & Zhang, W. (2020). Formation and immobilization of Cr (VI) species in long-term tannery waste contaminated soils. Environmental Science & Technology, 54(12), 7226-7235. DOI: 10.1021/acs.est.0c00156. 37 Graphical Abstract 38 3.1 Introduction Chromium (Cr) is a common contaminant due to its wide use in industrial processes such as metal plating, leather tanning and wood preservation.207 Total Cr concentrations in uncontaminated soils generally range from 2 – 200 mg Cr kg-1 soil globally,30 but anthropogenic activities increase soil Cr by several orders of magnitude, posing risks for human health and the environment.6 In soils, Cr is present in two common oxidation states which differ dramatically in terms of charge, chemical reactivity and toxicity.10 Cationic Cr(III) is immobile under most soil conditions due to adsorption on soil colloids and formation of insoluble complexes with hydroxides and organic ligands.93 In contrast, anionic Cr(VI) binds less extensively to soil colloids and thus has greater mobility.10 Cr(III) is an essential micronutrient for human health while Cr(VI) is a carcinogen.25,208 From an environmental risk perspective, remediation of Crcontaminated sites relies on the reduction of mobile and toxic Cr(VI) to less mobile and less toxic Cr(III).6 Consequently, knowledge of heavy metal speciation in soils is essential for risk assessment and remediation of contaminated sites.9 In soils, Cr(VI) is considered mobile and readily extracted. Spiking studies typically use a single extraction technique, reporting that available Cr(VI) represents 75 -100% of total added Cr(VI).209,210 However, actual contaminated soils have been reported to have a much lower fraction of available Cr(VI) than expected. Of the total Cr(VI) measured with X-ray adsorption near edge spectroscopy (XANES) analysis, only 0.5 ~ 10% was phosphate-exchangeable in soils collected from Cr-contaminated sites in Wisconsin (United States).211 In the previous work, soils from tannery waste sites had 12.6 ~ 42.5 mg kg-1 total Cr(VI) (as determined by alkaline digestion), with phosphate extractable Cr(VI) accounting for only ~10% of the total Cr(VI).128 Many studies attributed the low Cr(VI) availability in soils to the reduction of Cr(VI) to less 39 available Cr(III).94,212 But, significant amounts of total Cr(VI) in naturally contaminated soils suggest that Cr(VI) has not been reduced or leached out, but immobilized or stabilized in those soils.211,128 Diverse mechanisms may immobilize Cr(VI) in soils. Adsorption of Cr(VI) on Fe oxides in soils occurs rapidly at pH < 6.5, with the formation of outer-sphere (physical) or inner sphere (chemical) complexes.10,213 Given sufficient time, Cr(VI) could be incorporated into soil minerals structures via recrystallization process.214 At high concentration, Cr(VI) may precipitate as sparingly soluble salts such as PbCrO4, CaCrO4, and BaCrO4.215 These processes stabilize Cr(VI) and result in the lower Cr(VI) availability in soils.19 To the best of the knowledge, speciation of Cr in actual contaminated soils, and mechanisms responsible for Cr(VI) immobilization in longterm tannery waste contaminated soils is not fully understood. The objectives of the work were to (i) quantify the speciation of Cr in actual contaminated soils; and (ii) to elucidate the genesis of immobile Cr(VI) in such soils impacted by tannery waste. Specifically, (i) total concentrations, oxidation states and fractionation of Cr in contaminated soils were measured using chemical extraction techniques; and, (ii) synchrotronbased X-ray Fluorescence (XRF), XANES and theoretical modeling approaches were used for molecular characterization of Cr speciation and distribution in soils. 3.2 Materials and methods 3.2.1 Study area and soil sample collection The study area is located at Shuitou town, Pingyang county, Zhejiang province, in the southeast of People’s Republic of China. Over many years of continued leather tanning activities, the study area has been heavily impacted by Cr-containing tannery wastes.7 Due to the spatial 40 heterogeneity of Cr contamination, a hand-held X-ray fluorescence spectrometer (Genius 9000 XRF, Skyray Instruments USA) capable of in-situ determination of heavy metal concentrations was used to direct soil sampling. Composite surface soil (0-10 cm) samples at concentrations from 145 to 27,200 mg kg-1 were collected using a shovel from randomly selected locations in Shuitou-3 site (27.62327oN, 120.31337oE), which was reported as the most Cr-contaminated area.128,7 A composite sample was prepared from soils collected from 3 individual spots at each location. Three samples (S1-4, S2-1, S3-1) having the highest Cr concentration were used in this study for Cr speciation analysis. Specifically, soil 1-4 was collected from a tannery waste dumping site, not currently for agriculture uses. Soil 2-1 was gathered from farmlands that are currently being used by local inhabitants for agricultural purposes (such as growing vegetables) for both individual consumption and/or small commercial production for additional income. Soil 3-1 was also collected from farmlands that were growing vegetables but also used for leather airdrying. Tannery waste from the site was used as a reference of high levels of Cr. Once collected, soil samples and tannery waste were kept in polyethylene bags and stored in the dark at 4 oC to maintain field moist conditions. Field soils were used directly for Cr speciation analysis and portions of the samples were air-dried and ground to pass a 2-mm sieve prior to soil characterization. 3.2.2 Characterization of soil samples Soil pH was measured in 0.01 M CaCl2 and H2O with a 1:2 soil-to-solution ratio.216 Specifically, ten grams of air-dried soils was placed into a 100-mL beaker, added with 20 mL of 0.01 M CaCl2 solution or H2O. Then a glass rod was used to stir the suspension for 4 – 5 times during the next 30 minutes and then allow the suspension to settle for 30 minutes. The pH was recorded when the reading of the pH meter has stabilized (Thermo Scientific™ Orion 550A 41 benchtop pH/conductivity meter). Total C and N were determined by dry combustion using a total carbon/nitrogen analyzer (Costech 4010 Elemental Combustion System) at Northern Analytical Lab Services (NALS) of UNBC. The US EPA 9081 sodium acetate method was used to measure cation exchange capacity (CEC).217 Two grams of soil were mixed with 16.5 mL of 1.0 N sodium acetate solution, resulting in an exchange of the added sodium cations for the soil cations. The suspension was shaken and centrifuged, and the supernatant was decanted. The procedure was repeated to three times. Subsequently, the residue soil was washed with 16.5 mL of 99% isopropyl alcohol up to three times. Ammonium acetate solution (1.0 N) was then added, which replaces the adsorbed sodium with ammonium. The concentration of displaced sodium is in 50-mL solution then determined by inductively coupled plasma optical emission spectrometry (ICP-OES, Teledyne Leeman Prodigy) at NALS. CEC is expressed in terms of moles of positive charge absorbed per unit mass (cmol kg-1). Particle size analysis was determined using the pipette method.216 Ten grams of soil samples in the 600-mL breaker were added with 100 mL water and 10 mL of 1 M HCl slowly to remove the carbonates in soils. The suspension was left overnight in contact with HCl for a completed dissolution and the supernatant solution was siphoned off on the next day. After removal of carbonates, 10 mL H2O2 slowly added to samples to remove organic matter in the soil, with few drops of amyl alcohol can be added to suppress frothing. Volumes of H2O2 was added a few times until no more frothing occurred. Adding water to 400 mL, the beaker was placed on a hot plate and boil for about 1 hour to remove excess H2O2, and the supernatant liquid was siphoned off. Subsequently, water was added to make up volume about 400 mL and also 50 mL of Calgon solution (100 g L-1 of sodium hexametaphosphate at pH 8.3). To separate the sand in the sample, 42 soil suspension was carefully poured through a 300-mesh sieve (0.050-mm) which was placed above a 1-L sedimentation cylinder. Sand was retained on the sieve and weighed after drying overnight at 105oC (A, g), while silt and clay were washed through the sieve in the cylinder. After stirring the suspension, a closed pipet was carefully lowered to a depth of 10 cm to withdraw 20 mL aliquot in about 10 seconds. The aliquot was dried in the oven at 105 oC for 24 hours, which was the weight of silt and clay in the 20 mL suspension (b, g). Settling the suspension for 7 hours and 25 minutes (settling time for 0.002-mm clay at 22 oC according to the Stock’s law) (Figure 3.1), a closed pipet was used to withdraw 20 mL aliquot a depth of 10 cm in about 10 seconds. The aliquot was dried in the oven at 105 oC for 24 hours, which was the weight of clay in the 20 mL suspension (c, g). The clay in the sample was calculated as 𝐶(𝑔) = 𝑐 × 50 and silt in the samples was calculated as 𝐵(𝑔) = (𝑏 − 𝑐) × 50, respectively. Figure 3.1 Particle size analysis apparatus using the pipette method. 43 Total elemental contents of samples were analyzed using lithium metaborate fusion followed by nitric acid digestion resulting in a 2.4 M HNO3 solution. Elemental contents in the digests were determined using ICP-OES to quantify each element (British Columbia Ministry of Environment Analytical Laboratory Services, Victoria, British Columbia). Soil certified reference materials Till-1 and Till-3 were used for quality control of total elemental analysis. The mineralogical composition was determined by X-ray diffraction (XRD). XRD patterns were recorded between 0° and 90° (2θ) using a Rigaku MiniFlex X-ray diffractometer. Phase identification was performed using the Inorganic Crystal Structure Database (ICSD) under MDI Jade software. Thermogravimetric analysis was employed to determine CaCO3 in soils218 using a thermogravimetric analyzer (TGA) (Discovery TGA, TA instruments USA). A total of 15 – 20 mg of dry soil (A, mg) was positioned on the platinum sample plate and the weight loss of the sample heated from 30 oC to 900 oC was measured by a TGA analyzer. CaCO3 was decomposed to CO2 within the temperature range of 500 – 800 oC. The difference between the weights at 500 °C (B, mg) and at 800°C (C, mg) was attributed to the weight of generated CO2. Applying the conversion factor of 2.274, the ratio of the molecular weight of CaCO3 100.09 and that of CO2 44.01, CaCO3 was calculated as 𝐶𝑎𝐶𝑂3 % = 𝐵−𝐶 𝐴 × 2.274 × 100%. 3.2.3 Chemical extraction analysis of Cr speciation A detailed definition of metal speciation refers to (i) the identity of the element, (ii) oxidation states, (iii) physical states (i.e., a free ion or complexed to a ligand in solution, in the gaseous phases, or on solid phases), and (iv) the molecular geometry and coordination environment.9,219 44 In this study, all of the above components were identified by combining chemical extraction and synchrotron-based spectroscopy methods to better interpret Cr speciation and reactivity in soils. Total Cr(VI) analysis was conducted according to USEPA 3060A and USEPA 7196A.220,221 USEPA 3060A entails alkaline digestion to extract the soluble, absorbed and precipitated forms of Cr(VI) compounds from soils. USEPA 7196A determines Cr(VI) colormetrically by reaction with diphenylcarbazide (DPC). Specifically, a 50 mL alkaline solution of 0.28M Na2CO3/0.5M NaOH (pH > 11.5) was added to 2.5 g field moist soil in a 250-mL Erlenmeyer flask, with the addition of 400 mg of MgCl2 and 0.5 mL of 1.0M phosphate buffer. The mixture was heated to between 90 and 95 oC for 90 min prior to filtration. The filtered digest was adjusted to pH 7 – 8 by the addition of 5.0 M HNO3 and filtered with 0. 45 μm diluted to 100 mL with distilled water. After 2 hours, the solution was diluted 10-fold to 50 mL and adjusted to a pH of 1.5 – 2.5 with 10% (v/v) H2SO4. The Cr(VI) in solution was determined by reaction with diphenylcarbazide and measured by spectrophotometry at 540 nm. The accuracy of total Cr(VI) analysis was evaluated by analyzing a certified reference sample of soil CRM 041 (LRAB5366) (SigmaAldrich, Canada). Good agreement of the measured Cr(VI) content (114 ± 0.6 mg kg-1) was achieved with certified Cr(VI) content (130 mg kg-1, 13% standard deviation, 91 – 169 mg kg-1 acceptable interval). The recovery of K2Cr2O7 and PbCrO4 in CRM 041 was 99.80 ± 1.87% and 87.09 ± 1.21%, respectively. Available Cr(VI) in soils was measured using three separate single-extraction solutions: i) phosphate buffer209 (PBE; 5 mM K2HPO4 and 5 mM KH2PO4 at pH 7.2, in which the competing phosphate anion was concentrated enough to desorb Cr(VI) from exchange sites); ii) distilled water at pH 5.7;222 and iii) Synthetic Precipitation Leaching Procedure 223 solution (SPLP solution composed of a 40:60 weight percentage mixture of concentrated HNO3 and H2SO4 45 which was then diluted with distilled H2O to pH 4.20). Each separate extraction entailed adding 20 mL aliquots of extracting solution (PBE; water; SPLP) to 2.5 g of field moist soil (known moisture content). Tubes were shaken at 180 rpm for 30 minutes and centrifuged at 8,200 g-force (× g) for 15 minutes (JA-25.50 fixed-angle aluminum rotor, Beckman Coulter Ireland) to separate soil and the extractant. Supernatant was then filtered through a 0.45 µm membrane filter and Cr(VI) was determined colormetrically with DPC. A two-step sequential extraction method was developed for fractionation analysis of Cr(VI) in soils. In brief, 2.5 g of field soil samples was placed in 50 ml centrifuge tubes and 20 ml of phosphate buffer was added. The phosphate buffer extraction (PBE) process as described above, represented available Cr(VI). The residue soil in the tube was transferred to 250 mL Erlenmeyer flasks for alkaline digestion (USEPA 3060A) described as above, designated the immobile Cr(VI). James & Bartlett used the sum of three times of PBE Cr(VI) as exchangeable or available Cr(VI) in spiking studies.209 In this case, only a single PBE represented the available Cr(VI), since (i) very small amount of Cr(VI) (< 1 mg kg-1 for all three samples) could be extracted by the second and third PBE, and (ii) measured immobile Cr(VI) content declined with multiple times of PBE (Figure 3.2). 46 Figure 3.2 Sequential extraction results for Cr(VI) fractions (mg kg-1) in soil 1-4, soil 2-1, soil 31 and certified reference soil sample CRM041. Numbers (1, 2, 3) refers to the repeated times of phosphate buffer extraction (PBE). Column 1 – single PBE; Column 2 – two times of PBE; Column 3 – three times of PBE. The content of measured immobile Cr(VI) is marked (green bar). Mean (and standard error) of calculated total Cr(VI) (mg kg-1) is shown in the table. Different letters (i.e., a, b, c) are significantly different (p < 0.05) in calculated total Cr(VI) contents by analysis of variance (ANOVA). 47 3.2.4 SEM-EDXS The morphology and semi-quantitative in situ elemental composition of selected points of interest were observed using a Philips XLS 30 SEM equipped with an energy dispersive x-ray system (EDS) at NALS. 3.2.5 XANES experiments Cr K-edge (5989 eV) x-ray adsorption near edge adsorption (XANES) spectra were collected at the Canadian Light Source (CLS) using the 06ID-1 Hard X-ray Micro-Analysis Beamline (HXMA). A Si (111) double crystal monochromator was calibrated at Cr K edge at 5989 eV by using the metallic Cr foil provided by Exafs Materials, with an energy revolution (ΔE/E) of 10-4. Cr standard spectra for Cr (VI) and Cr (III) reference solids were recorded in transmission mode. XRF maps and spectra for Cr-contaminated soils were recorded in fluorescence mode using a Canberra 32 element Ge solid state detector. Soil samples (bulk soil, before and after PBE extraction) were prepared and sealed with Kapton tape. Several reference compounds were used: Cr(III) acetylacetonate [Cr(C5H7O2)3], Cr(III) chloride hexahydrate (CrCl3 · 6H2O), Cr(III) sulfate [Cr2(SO4)3], lead chromate (PbCrO4), calcium chromate (CaCrO4) and potassium dichromate (K2Cr2O7). For each sample, two scans were sufficient to obtain a good signal/noise ratio. XANES data processing and analysis was conducted using the Athena software from Demeter,224 including energy calibration, pre-edge background subtraction and post-edge normalization. 3.2.6 FDMNES Theoretical analysis of experimental Cr K-edge XANES spectra was performed using FDMNES (Finite difference modelling of the near-edge structure) code.225,226 In this study, Cr K- 48 edge XANES of Cr(III) phases with the increasing cluster radius were calculated by FDMNES, in order to identify the optimum cluster radius size from simulation that fit with experimental data. Calculated XANES spectra served as the basis for the interpretation of the experimental data. 3.2.7 LCF analysis Quantitative Cr speciation in soils was determined by linear combination fitting (LCF)32,228 with contributions of selected standards spectra from theoretical modelling and experimental data, respectively. The Chi-square (χ2) and residue factor (R-factor) were used for primary goodness of fit assessment, while energy shift (ΔE0) was also considered.227,229 3.3 Results 3.3.1 Soil properties Selected chemical and physical properties of soils and tannery waste are shown in Table 3.1. The pH of the tannery waste and soil samples was high as measured in both H2O and CaCl2, ranging from 7.40 (S3-1) to 7.96 (S2-1). Tannery waste contained up to 18.0% CaCO3 (Figure 3.2), accounting for 51.3% of total C (4.2%). Compared to tannery waste, soil samples had a lower proportion of total C in the form of CaCO3, ranging from 29.2% (S1-4) to 37.7% (S3-1) (Figure 3.3). Total N content (%) in tannery waste and soils ranged from 0.16 to 0.28 and CEC (cmol kg-1) in soils ranged from 22.9 to 39.9. The sand fraction accounted for more than 50% of the soil particle size distribution, with clay fraction approximately 20%. According to the textural triangle,216 soil samples in this study were classified as sandy clay loam. 49 Table 3.1 Selected physical and chemical properties (means with standard deviation, n=3) of contaminated soils and tannery waste used in this study. Properties Tannery waste Soil samples S1-4 S2-1 S3-1 pH (H2O) 7.87 (0.08) 7.75 (0.05) 7.96 (0.02) 7.63 (<0.01) pH (CaCl2) 7.77 (0.05) 7.59 (0.02) 7.50 (0.02) 7.40 (0.05) CEC (cmol kg-1) ND 39.9 (0.2) 22.9 (0.4) 28.2 (1.2) Total C (%) 4.20 (0.20) 6.89 (0.18) 2.42 (0.03) 3.76 (0.04) Total N (%) 0.18 (0.05) 0.25 (0.05) 0.16 (0.01) 0.28 (0.02) CaCO3 (%) 17.97 16.79 6.39 11.81 Sand (%) ND 62.2 (1.3) 54.8 (1.4) 58.2 (1.4) Silt (%) ND 16.4 (1.3) 22.0 (1.1) 19.1 (1.7) Clay (%) ND 21.4 (1.8) 23.2 (0.8) 22.7 (1.3) Texture Sandy clay loam Total elemental analysis Si (wt%) ND 18.6 (0.3) 27.1 (0.3) 25.0 (0.2) Fe (wt%) 3.00 (0.76) 4.82 (0.02) 3.22 (0.12) 2.87 (0.03) Al (wt%) 2.31 (0.26) 4.40 (0.05) 5.68 (0.04) 5.55 (0.02) Ca (wt%) 10.70 (1.05) 6.65 (0.18) 2.33 (0.18) 3.34 (0.09) Mg (wt%) 2.28 (0.09) 1.71 (0.02) 0.48 (0.03) 0.45 (0.02) Ni (mg kg-1) 75 (3) 69.1 (3.6) 31.9 (5.4) 37.2 (8.5) Co (mg kg-1) 35.1 (6) 41.6 (5.1) 30.8 (6.6) 19.7 (2.0) P (mg kg-1) 0.12 (<0.01) 0.15 (<0.01) 0.07 (<0.01) 0.11 (<0.01) S (mg kg-1) 0.73 (0.05) 0.20 (0.00) 0.08 (0.00) 0.09 (0.02) Cr (mg kg-1) 52370 (257) 27200 (900) 6479 (815) 4317 (4) Ti (mg kg-1) ND 3885 (54) 3808 (45) 3756 (27) Mn (mg kg-1) 605 (56) 830 (15) 757 (28) 747 (4) Ba (mg kg-1) 347 (14) 572 (30) 570 (13) 618 (4) ND = not determined; wt %= weight percentage. 50 Figure 3.3 CaCO3 content in soils (S1-3, S2-1 and S1-4) and tannery waste as determined by thermogravimetric analysis (TGA) 51 Total elemental contents of soils by the fusion method) and by the sludge sample (acid digestion) is shown in Table 3.2. A more complete dissolution of Cr-contaminated soils was observed by lithium metaborate fusion method than by the strong acid (HNO3 & HClO4) microwave digestion method. This is a result of incomplete digestion of silicates by the acid digestion method, resulting in only a partial extraction of some metals such as Al, K, Mg (personal communication with BC Ministry Analytical Lab, Victoria, British Columbia). Relatively high total Cr was observed in all samples with concentrations ranging from 4320 – 27200 mg kg-1. Concentrations of other elements (%) in soils ranged widely for Si (18.6 – 25.0), Ca (2.33 – 6.65), Fe (2.87 – 4.82), Al (4.40 – 5.68), K (1.24 – 2.35) and Mg (0.45 – 1.71). Compared to soils, tannery waste contained higher concentrations of Ca (10.70%), S (0.73%) and Cr (52.4 g kg-1). Other elements, such as Ba and Mn content, occur at much lower concentrations (mg kg-1) varying from 605 to 830 and from 347 to 618, respectively. 52 Table 3.2 Mean (and standard error) (n = 3) of elemental contents in soils collected around three leather tanning sites in Shuitou (southern China). S 1-4 S 2-1 Acid digest Fusion Acid digest Fusion Elements (Si – S:%; Cu – B: mg kg-1) Si N/A 18.6 (0.3) N/A 27.1 (0.3) Al 2.44 (0.06) 4.40 (0.05) 2.89 (0.07) 5.68 (0.04) Ca 5.92 (0.17) 6.65 (0.18) 1.73 (0.06) 2.33 (0.18) Fe 4.05 (0.26) 4.82 (0.02) 2.38 (0.06) 3.22 (0.12) K 0.18 (0.01) 1.24 (0.01) 0.37 (0.02) 2.33 (0.03) Mg 0.89 (0.01) 1.71 (0.02) 0.30 (0.01) 0.48 (0.03) Na 0.18 (0.33) 0.54 (0.17) 0.36 (0.09) 0.58 (0.12) P 0.12 (<0.01) 0.15 (<0.01) 0.06 (<0.01) 0.07 (<0.01) S 0.16 (<0.01) 0.20 (<0.01) 0.05 (<0.01) 0.08 (<0.01) Cu 69.1 (15.7) 70.0 (4.5) 26.3 (1.9) 32.2 (3.3) Zn 379 (14) 439 (20) 212 (15) 214 (12) As 8.35 (0.08) N/A 7.45 (0.05) N/A Ba 395 (14) 572 (30) 186 (2) 570 (13) Cd 0.87 (0.46) N/A 0.36 (0.01) N/A Co 36.3 (0.7) 41.6 (5.1) 30.7 (5.6) 30.8 (6.6) Cr 2471 (45) 27200 (900) 2278 (108) 6479 (815) Ni 37.4 (3.1) 69.1 (3.6) 13.1 (0.4) 31.9 (5.4) Mo 4.47 (0.7) 7.5 in all cases, with significant CaCO3 present. In contrast, non-contaminated soils in Shuitou town had total Ca content of 0.43% (equivalent to 1.08% of CaCO3) and a pH of 6.1.3 The high pH and CaCO3 contents were mainly attributed to the contamination by tannery waste that had pH of 7.87 and a CaCO3 content of 18.0%. Alkaline conditions are consistent with the high concentration of Cr(VI) in these soils. Main electron sources for Cr(VI) reduction in soils include organic matter and Fe(II) minerals.10 Under alkaline conditions, anionic Cr(VI) adsorption onto organic matter and Fe oxides is limited, thereby decreasing the rate of Cr(VI) reduction.212 Compared to acidic (pH=1) and neutral condition (pH =7), Cr(VI) reduction by magnetite at high pH conditions (pH =13) was limited to < 20% reduction capacity.241 3.4.2 Cr distribution and oxidation states in soils Although the absolute concentration of Cr(VI) was high, 97 % or more of Cr was present as Cr(III) species (Table 3.1). The low proportion of Cr(VI) compared to Cr(III) was expected because Cr(III) salts are used in leather tanning and oxidation of Cr(III) to Cr(VI) is independently limited by water-soluble Cr(III) concentration, pH, initial available surface area and ionic strength.84 Recent studies have shown that the addition of 1 mM solution of CrCl3 to a MnO2-treated soil did not result in complete oxidation. Rather, only 4.7% of the added Cr(III) 67 generated Cr(VI).242 The oxidation of Cr(III) by MnO2 was gradually inhibited by the formation of Cr(OH)3 coverage of MnO2 surface.243,90 At the sub-millimeter scale, Cr was heterogeneously distributed and accumulated in a few hot spots embedded within a matrix of Mn- and Fe-oxides in these long-term contaminated soils (Figure 3.5). Direct XANES investigations showed the coexistence of Cr(III) and Cr(VI) species in some Cr hot spots, but other hot spots contained only Cr(III) species (Figure 3.6b). The occurrence of Cr(VI) in Cr hot spots is attributed to Cr(III) oxidation by Mn oxides, since a close spatial relationship between Cr and Mn was observed in Cr hot spots (Figure 3.5). Haulasen and Fendorf (2017)244 demonstrated Cr(VI) generation was localized with Mn hotspots. Faundear et al. (2009)245 reported the occurrence of Cr(VI) (up to 20% of total chromium mass) corresponded with locations of Mn-rich area (up to 21.7% MnO by mass) in highly weathered lateritic soils, with almost no Cr(VI) detected elsewhere. In most natural environments, Mn(III/IV) oxides are the only known oxidant capable of oxidizing Cr(III) to Cr(VI).8,245 Those Cr hot spots containing only Cr(III) may reflect the limited Cr(III) oxidation by Mn oxides, and/or the fast reduction of Cr(VI) to Cr(III) by Fe. Scanning electron microscope (SEM) images and elemental spectra detected in EDS revealed Fe (2.77 – 17.8 mass%), Cr (0.51 – 27.23 mass%) and Mn (0 – 1.4 mass%) on random selected spots in soil 3-1 (Figure 3.12 and Table 3.5). Positive and significant relationships were observed in mass % for Fe & Cr (R2 = 0.726, p < 0.001), which was in agreement with previous studies that Fe oxides were strong storage sinks for Cr.7,246 Reduction of Cr(VI) to Cr(III) indirectly by Fe(III) hydr(oxides) such as hematite and goethite could be via a closely coupled, biotic-abiotic reductive pathway under reducing environment.247 Wielinga et al. (2001)247 reported that Fe(II) generated via Fe(III) reduction by iron-reducing bacteria Shewanella alga strain BrY effectively reduce Cr(VI) to Cr(III). Reducing 68 environment in soils usually occurs at anaerobic zones in the center of soil aggregates,248 and could be created by frequent flooding in the study area.7 In addition, Fe(III) hydr(oxides) such as goethite may have structural Fe(II) in the solid phase that was reactive for Cr(VI) reduction.249 Therefore, if Cr(VI) was generated in these Cr spots with high Fe contents, it was likely to undergo reduction at iron-reducing condition, since Fe(II) is one of the main reducing agent for converting Cr(VI) to Cr(III) in soils.250,251,252 3.4.3 Cr availability in soils Availability of heavy metals in soils includes the fraction dissolved in the pore water and the fraction potentially desorbed from the soil solids, governed by soil properties and soil processes (e.g. adsorption/desorption, dissolution/precipitation).92 Cr(III) has very low availability in soils, because Cr(III) hydroxides form easily at pH 4 – 9 and the retention of Cr(III) on soil minerals is favorable under most conditions.10 Our results in which only 2.0 to 18 mg kg-1 of Cr(III) (Table 3.3) was water-extractible from contaminated soils (4317 – 27200 mg Cr kg-1 soil-1) demonstrate the low availability of Cr(III) in such soils. Similarly, Kamaludeen et al.253 reported only 1.9 mg kg-1 water-soluble Cr(III) in soils contaminated with 102.0 g Cr kg-1. In contrast, Cr(VI) is highly mobile or available in soils.10 Twelve Australia soils were spiked with 0 - 500 mg L-1 Cr(VI) and allowed to equilibrium for 16 hours; subsequently, more than 90% of added Cr(VI) was desorbed after 2-h equilibrate with 1M KNO3.210 In soil developed from ultramafic rocks, Cr(VI) generated on MnO2 was extractable by phosphate solution indicated by XANES analysis.240 However, in the present study, total Cr(VI) content in soils was as high as 144 mg kg-1, and > 90% of Cr(VI) was in the immobile fraction indicated by both chemical extraction methods and XANES analysis (Table 3.3; Figure 3.6). These results do not support the hypothesis that all Cr(VI) was highly mobile and easily extractable, but were consistent with 69 many studies in naturally Cr-contaminated sites211,215 Chrysochoou et al.(2010)215 concluded that in those Cr(VI) contaminated soils with high Pb content of 15.0 g kg-1, Cr(VI) precipitates as insoluble PbCrO4 which significantly lower the mobility of Cr(VI). Fandeur et al. (2009)245 reported that Cr(VI) was generated at the boundary of Mn and Fe oxides in studied lateritic regolith samples and 50% of the Cr(VI) was not extractable due to the reabsorption of Cr(VI) onto surrounding Fe-oxyhydroxides after oxidation by Mn oxides. The recrystallization depicted by Zhu et al. (2019)214 showed that when ferrihydrite transformed to more stable hematite and goethite in the presence of silicates with aging, Cr(VI) on ferrihydrite was incorporated into the hematite and goethite structure and became immobilized. In our case, all of those reactions are possible. High Ca content in soils (2.33 – 6.65 % ) (Table 3.1) and the significant relationship of Ca and Cr (R2 =0.629, p = 0.004) in mass % from EDS (Figure 3.12) suggested that Cr(VI) may precipitate as moderately soluble CaCrO4 (Ksp 7.1 x 10-4) 254. Significant association of Cr and Fe (R2 = 0.726, p < 0.001) in mass % from EDS (Figure 3.12) and the distribution of Cr that embedded with Fe oxides (Figure 3.5) favor a reabsorption or slow recrystallization of Cr(VI) onto Fe phases, reducing Cr(VI) availability in soils. 3.4.4 Cr speciation in soils Beyond the determination of oxidation states and mobility, FDMNES modelling identified CrFeO3 (space group R3̅) as the dominant Cr(III) species and suggested CrOOH (space group Pbnm) was the second Cr(III) species in soil 2-1 in this study (Figure 3.8; Figure 3.9). The geometric structure of the two Cr(III) species is presented in Figure 3.13. Although fresh tannery waste contained the main Cr2O3 phase from XRD result (Figure 3.4), Cr2O3 was not the main phase in soil 2-1 identified by FDMNES modelling (Figure 3.10), indicating the transformation from Cr2O3 to CrFeO3 in soils at long-term scale upon weathering. The formation of CrFeO3 was 70 likely via the coherent slow intergrowth of Fe2O3 (hematite) and Cr2O3 (eskolaite).233,255 Occurrence of CrFeO3 (R3̅) is highly possible in the present study, since Cr2O3 was introduced into soils by tannery waste contamination and our previous micro XRD analysis showed Fe-rich minerals from Shuitou soils were predominantly α-Fe2O3 (hematite) and α-FeOOH (goethite).128 The formation of CrOOH (Pbmn)256 may result from Cr3+ substitution onto Fe oxides and/or Cr(VI) reduction. Isovalent Cr3+ occupied the octahedral positions within the oxygen framework in α-FeOOH leading to isostructural product α-CrOOH (beacewellite).257,258 Reduction of Cr(VI) by magnetite could form α-CrOOH polymorph.259 Figure 3.13 Crystal structures and electron density maps of (a) CrFeO3 (R3̅) and (b) CrOOH (Pbnm). Selected planes (i, ii, iii, iv) are through the Cr-O-Fe and Cr-O-Cr ions labeled in the crystal structures. The contour unit is e/Bohr3 (1 Bohr = 0.5239 Å). 71 3.4.5 Genesis of immobile Cr(VI) in long-term contaminated soils Although fresh tannery sludge generally contained little or no Cr(VI), tannery waste contaminated sites had toxic quantities of Cr(VI) at various percentage of total Cr.211,128,260 As discussed in the previous section, the genesis of Cr(VI) in contaminated soils was via Cr(III) oxidation by Mn oxides. Various Cr(III) species including dissolved Cr(III) ions, iron-based CrxFe1-x(OH)3 and even the most stable product Cr(OH)3 could be oxidized by MnO2.243,90,261 In natural environment, Cr(III) oxidation expected to entail two steps: (i) formation of Mn(III/IV) through microbial or chemical Mn(II) oxidation and (ii) Cr(III) reaction with Mn(III/IV) oxides that result in Cr(VI) genesis.262,263,264,265 In this way, Mn was recycled and only small amounts of Mn may have long-term impacts on biogeochemical processes of Cr(III) oxidation in contaminated soils. In the present study, both CrFeO3 and CrOOH appear to be involved in conversions to Cr(VI) species. Component weight collection analysis of LCF results was displayed in Figure 3.14. We observed a negative correlation between CrFeO3 and CaCrO4 in phase (i) and (iii), suggesting CrFeO3 serves as the electron donor for Cr(VI) production. Specifically, the weight percentage of CrFeO3 in 2-1-2 was 9.7% lower than in PBE-3. The decreased weight of CrFeO3 was mainly attributed to the oxidation to CaCrO4 (8%), with 1.7% converted to CrOOH. In phase (ii), the increase of CaCrO4 was positively corrected with CrFeO3, but negatively related to CrOOH, indicating CrOOH was oxidized to CaCrO4. The weight percentage of CrOOH drops from 29.5 (2-1-2) to 22.2 (PBE-4) which allows 2.9% of CrOOH to be converted to CaCrO4, with 4.2 % being transformed to CrFeO3. The oxidation mechanisms were further examined by density functional theory calculations (DFT) using CP2K program in the Appendix A.266 Deformation charge density of CrFeO3 and 72 CrOOH with δ-MnO2 showed that both CrFeO3 and CrOOH molecules lose electrons (corresponding to the large blue region) and are potentially oxidized to CrO42-, while Mn atoms receive electrons (corresponding the yellow region) and will be reduced (Figure 3.15). These results provide a theoretical basis to expect Cr(VI) formation in Shuitou soils by CrFeO3 and CrOOH localized at Mn hotspots. Figure 3.14 Component weight correlation among CrFeO3, CrOOH and CaCrO4 from LCF result. η refers to the conversion between components. 𝜂1→3 red arrow: CrFeO3 → CaCrO4; 𝜂2→3 blue arrow: CrOOH → CaCrO4; 𝜂1→2 green arrow: CrFeO3 → CrOOH; 𝜂2→1 green arrow: CrOOH → CrFeO3. For each of the point, the sum of weight of the three Cr species is equal to 1. 73 Figure 3.15 Deformation charge density of (a) CrFeO3 and (b) CrOOH with δ-MnO2. Isosurface values = ± 0.006 e Å–3, where yellow and blue represent electron-rich and electron-deficient regions, respectively. Mn: pink balls; O: red balls. After Cr(VI) is formed in soil environments, three fates await it: i) it will remain mobile; ii) it will become immobile via adsorption or precipitation; or, iii) it be reduced to Cr(III). Bartlett & James (1979)267 reported that more than 60% of total added Cr(III) (520 mg kg-1) to soils was oxidized to Cr(VI) within 24 hours and soluble Cr(VI) decreased to 24% total added Cr(III) after five months. They stated that Cr(VI) in soils was reduced or leached out.267 In clay soils amended with tannery sludge, 1.1% of total added Cr was oxidized to Cr(VI) during the first 5 months and Cr(VI) deceased from 27 to 5 mg kg-1 by the end of the 2 year experiment.265 Sequential extraction analysis indicated that soluble Cr and Cr(VI) were redistributed to more sparingly soluble fractions.265 In our case, total Cr(VI) content reached 144 mg kg-1 in long term contaminated soils impacted by tannery waste (Table 3.3), most of which was in forms inaccessible to solutions. As discussed previously, we argue that CaCrO4 precipitation is the main immobilization mechanism in soils with high Ca content (2.33 – 6.65%). This was further 74 confirmed by LCF analysis of XANES spectra with statistically acceptable fits (Figure 3.6). The strong association of Cr and Fe from EDS (Figure 3.12) suggested the reabsorption or recrystallization of Cr(VI) with Fe phases could be the second mechanisms for Cr(VI) immobilization in soils. 3.5 Summary The results in this study illustrate the genesis of immobile Cr(VI) fractions in these soils (Figure 3.16). Over large time scale after tannery waste contamination, Cr exists in soils mainly as Cr(III) species (CrFeO3 and CrOOH), but with environmentally high amounts of Cr(VI) present. XRF analysis revealed the strong spatial association of Cr with Mn, strongly suggesting that Mn oxides could oxidize Cr(III) into Cr(VI). After oxidation, Cr(VI) is further immobilized via precipitation as CaCrO4 and via readsorption or recrystallization with Fe oxides, indicated by EDS results and LCF analysis. Ultimately, these low total carbon, alkaline, sandy clay loam soils protect generated Cr(VI) species from both reduction and solubilization. It will be necessary to solubilize the immobile Cr(VI) species in such aged contaminated soils for reduction treatment, which is consistent with reports that insoluble Cr(VI) compounds in soils were recalcitrant to reduction treatment.215 Also, it is expected that remediation strategies of Crcontaminated sites may be improved by considering not only the reduction efficiency of toxic Cr(VI) to nontoxic Cr(III), but also the removal efficiency of total Cr to minimize the risk from re-oxidation of Cr(III) to Cr(VI). For example, Fe(II)-Al(III) layered double hydroxide (Fe-ALLDH) has been reported to be an efficient absorbent for Cr(VI). The redistribution of Cr(VI) into the Fe-Al-LDH structures protected immobilized Cr being solubilized or reoxidized to Cr(VI) in Cr(VI) contaminated soils.268 75 Figure 3.16 Simplified schematic of immobile Cr(VI) species formation in soils impacted by tannery sludge. 76 Chapter 4 AGING SHAPES CR(VI) SPECIATION IN SOILS: A 240-DAY INCUBATION EXPERIMENT Abstract To make sound decisions regarding management of heavy metal contamination in soils, it is necessary to understand contaminant transformations over extended periods. In this study, sequential extraction methods were applied to quantify the changes of Cr fractions [available Cr(VI), immobile Cr(VI) and immobile Cr(III)] in five contrasting soils spiked with Cr(VI) over a 240-day incubation. Results showed that available Cr(VI) in soils continually decreased during aging, with a sharp decline occurring in the first 30 days. The best fit of available Cr(VI) data was obtained using an Elovich model for Brunisol and Anthrosol-1, a fractional power model for Anthrosol-2, and a pseudo first-order kinetic model for Luvisol-1 and Luvisol-2. After aging for 240 days, immobile Cr(VI) increased by 4.5 – 31% and immobile Cr(III) increased by 68 – 95% of total spiked Cr(VI) in Brunisol, Anthrosol-1 and Anthrosol-2. The two Luvisol soils had relatively high reduction rates with no Cr(VI) immobilized. A multi-reaction model was developed in MATLAB Simulink toolbox to describe transformation flow rates among soluble Cr(VI), adsorbed Cr(VI), immobilized Cr(VI) and immobilized Cr(III) in soils with aging. It is concluded that (i) Cr(VI) reduction and immobilization were occurring concurrently in soils and competing for available Cr(VI) species; (ii) Cr(VI) reduction is favored by low soil pH and high organic carbon, while Cr(VI) immobilization occurs with cations (such as Ca2+) and Fe oxides. This chapter has been published as Shi, J., McGill, W. B., Rutherford, P. M., Whitcombe, T. W., & Zhang, W. (2022). Aging shapes Cr (VI) speciation in five different soils. Science of The Total Environment, 804, 150066. DOI: 10.1016/j.scitotenv.2021.150066. 77 Graphical Abstract 78 4.1 Introduction Contamination of soils by Cr(VI) is caused by a variety of anthropogenic activities including mining, metallurgy, electroplating, pigments production, tanning and wood preservation.269 Naturally occurring Cr(VI) derived from the weathering of ultramafic soils has also been reported.33,240 Chromium is present in two oxidation states in soils and the environment, Cr(III) and Cr(VI).10 Cationic Cr(III) is immobile under most soils conditions,93 and Cr(III) is an essential micronutrient for human health.208 In contrast, anionic Cr(VI) is highly soluble, which may leach into and contaminate groundwater. The oxyanionic character of Cr(VI) also allows it to mimic sulfate or phosphate in passing through cellular membranes, resulting in Cr accumulation in vegetables grown on contaminated sites.8 In addition, Cr(VI) is toxic to plants and animals because of its carcinogenic and mutagenic properties.270 The hazards imposed by Cr(VI) species motivate inquiry into its fate in soils. After entering into soils, Cr(VI) adsorption occurs rapidly on positively charged colloids, such as Fe oxides, at low soil pH.271 These complexes have been reported to be reversible with a change in pH; and competition by both sulfate and dissolved inorganic carbon for adsorption sites suppressed binding.272 Cr(VI) reduction to Cr(III) is common in soils. Cr(VI) reduction by aqueous Fe(II) can be rapid with complete reduction occurring within a few minutes.97 The rate of Cr(VI) reduction by humic acids has been shown to be on the order of days at pH 2 – 5.12 Reduction of soluble Cr(VI) by soils has been found to be first-order with respect to Cr(VI).15,264 These studies illustrate the fate of Cr(VI) in soils over relatively short reaction times on the order of hours to days. However, transformation of trace metals is likely to continue over time spans ranging from months to years.273 Continuous slow reactions that decrease the availability of contaminants in soils have been defined as the aging process.19 79 Previous studies suggested that the aging processes include substitution for a matrix ion, entrainment of the ion into the solid phase, surface precipitation, and diffusion into micropores. 98 These processes may lead to more extensive transformations of metals in soils. Liang et al. (2014) reported that spiked arsenic in soils was transformed from an available form and incorporated into amorphous and crystallized Fe/Al fractions bonded to arsenic after 33 weeks.20 Wang et al. (2017) reported that after a one-year incubation period, soluble selenium accumulated in the residual fraction in acidic soils.21 The previous work has demonstrated the complex speciation of Cr in long term Cr-contaminated soils in which the majority of Cr(VI) was immobile.274 These findings suggest continuous transformations of trace metals in soils during aging. However, the transformations of Cr(VI) species in soils during aging are still not fully understood. Understanding time-dependent Cr(VI) transformations in soils is necessary for risk assessment and remediation of Cr-contaminated sites. The objectives of this study were (i) to determine the transformation kinetics of Cr fractions in soils with increasing retention time, (ii) to investigate the impact of soil properties on Cr(VI) transformations with respect to both oxidation/reduction and solubilization/immobilization during the aging process, and (iii) to establish a model for describing the change of various Cr fractions in soils with aging. 4.2 Materials and methods 4.2.1 Soil sampling Five contrasting soils from three sites were used to meet these objectives. Two of the soil samples used in this study were collected from the reported contaminated site in Shuitou town, Zhejiang Province, China (27o 37’ 24” N, 120o 18’ 48” E).7,274 Soils in this site were impacted by long-term tannery activities, resulting in a pH > 7 and high concentrations of Ca2+, which were 80 expected to favor Cr(VI) immobilization during aging. Surface soils (0 – 15 cm) were collected from a non-growing area (Anthrosol-1) and a bokchoy growing area (Anthrosol-2). Anthrosol is defined according to the Chinese System of Soil Classification.275 Two contrasting soils were selected for comparison (Luvisol-1 & Luvisol-2) with low soil pH and higher organic carbon content. The Luvisolic soil samples were obtained from agricultural farmlands approximately 10 km north of Prince George, British Columbia, Canada (54o 04’ 24” N, 122o 48’ 02” W) (Dawson, 1989).276 Surface soils (0 – 15 cm) were collected from a fallow area (Luvisol-1) and from a hay growing area (Luvisol-2) which had recently been converted from forest to cropland. These acidic Luvisol soils were not expected to be favorable for Cr(VI) immobilization but to support Cr(VI) reduction in the presence of organic carbon.15,71 The last soil sample consisted of Brunisolic B horizon material from Bear Lake, British Columbia, Canada (54 o 29’ 41” N, 122 o 41’ 02” W). Brunisolic and Luvisolic are soil orders in the Canadian System of Soil Classification.277 A Brunisolic soil (sandy and low organic carbon), was expected to support slow Cr transformations compared with the other four soils. Consequently, a combination of the five soil samples used in this study provided a wide range of soil pH, soil organic matter content and cation concentrations, which were expected to exhibit different kinetics of Cr transformations with aging. All soil samples were air-dried, homogenized, and ground to < 2 mm for analysis and the aging incubation. 4.2.2 Characterization of soil samples Soil pH was measured in aqueous 0.01 M CaCl2 and H2O with a 1:2 soil-to-solution ratio.216 Total C and N were determined by dry combustion using a total carbon/nitrogen analyzer. Thermogravimetric analysis was used for measuring inorganic carbon contents (SIC) in bulk soils.278,218 Cation exchange capacity (CEC) was measured according to the US EPA 9081 81 sodium acetate method.217. Particle size analysis was performed using the pipette method.216 The details of these methods have been described in section 3.2.2. Total elemental contents of samples were analyzed using acid (HNO3 & HClO4) block digestion (SCP Science DigiPREP MS). Elemental contents in the digests were determined using inductively coupled plasma optical emission spectrometry (Agilent Technologies 5100; ICP-OES) to quantify each element. The mineralogical composition was determined by X-ray diffraction (XRD). XRD patterns were recorded between 3° and 90° (2θ) at a rate of 10 °/min and a resolution of 0.02°/step using a Rigaku MiniFlex X-ray diffractometer with Cu-Kα at 40kV and 15 mA. Water-soluble Cr and Ca was extracted from a mixture of dry soil with deionized water (1:10 w/v), shaken for 1 hour and centrifuged at 8200g for 15 minutes at 20oC. The extracts were analyzed using ICP-OES. Total porosity of soils was measured using tension and pressure techniques.279,280 Air-dried soil samples were gently compacted and filled in buffer rings, to have intact soil cores with 55 mm in diameter and 10 mm in height (Vb , cm3). The cores enclosing the sample were placed on the ceramic plate and wetted from below with distilled water. After 24-hour saturation, soil columns were placed in the pressure plate extractor and left until equilibrated at the desired pressure. Field capacity is equivalent to a moisture tension of 0.10 bar for sandy soils and to 0.33 bar for heavy textured soils, and permanent wilting point is equivalent to a moisture tension of 15 bar. Two types of pressure plate extractors were used (1600G1 5 – bar and 1500G1 15 – bar pressure plate extractors; Soilmoisture Equipment Crop., 2001), depending on the pressure applied. After 5 – 7 days of equilibrium time, wet soils in cores were taken out and weighed (mw + md , g), and the weight of dry soils were determined after drying the sample (𝑚𝑑 , g) in an oven 82 at 105oC for 24 hours. Bulk density (Db , g cm-3) and total porosity (TP, %) was calculated as follows: Db = TP = (1 − md ⁄V b Db ⁄D ) × 100% p Where Db = dry soil bulk density (g cm-3); md = the weight of oven-dried soil (g); Vb = the volume of the soil column (cm3); TP = total porosity (cm3 of pore volume (cm3 soil)-1); Dp = soil particle density (g cm-3), taken to be 2.65; mw = the mass of water in soils (g), calculated by the difference of soil mass before and after over-drying. 4.2.3 Soil incubation Air-dried soils (2.0 g oven dry equivalent) were positioned in 50-mL polypropylene centrifuge tubes. A volume of 0.3 – 0.5 ml K2Cr2O7 solution was added to each soil unit with a pipette, producing soil units amended with 15 – 500 mg Cr (VI) kg-1. Deionized water was then added to raise the soil moisture content to 60% of total porosity. The experimental design entailed fifteen treatments (5 soils × 3 concentrations) with six replicates (Table 4.1). To minimize the water loss from the soil, the tubes were placed in 2-Liter Mason jars containing enough water to maintain humidity. The sample tubes were stored uncapped, and the jar was covered by a lid with a 2-mm opening to allow for gas exchange. Jars were placed in a growth chamber in the dark (EGC M Series Growth Chambers) at a constant temperature of 25oC. The mass of tube and sample was monitored every seven days for the entire aging incubation and water was added to compensate for evaporation loss. At increasing aging periods (1, 2, 4, 8, 15, 30, 60, 90, 120, 180, and 240 days), soil samples in centrifuge tubes were directly used for Cr analysis. 83 Table 4.1 Summary of spiking treatments (n = 6)*. Treatment Soils Spiked with Cr(VI) £ Spiked con. 0.30 ml x 500 mg L-1 75.00 0.30 ml x 750 mg L-1 112.50 3 0.30 ml x 1000 mg L-1 150.00 4 0.50 ml x 750 mg L-1 187.50 0.50 ml x 1000 mg L-1 250.00 6 0.50 ml x 1500 mg L-1 375.00 7 0.50 ml x 1000 mg L-1 250.00 0.50 ml x 1500 mg L-1 375.00 9 0.50 ml x 2000 mg L-1 500.00 10 0.30 ml x 100 mg L-1 15.00 0.30 ml x 200 mg L-1 30.00 12 0.30 ml x 500 mg L-1 75.00 13 0.30 ml x 200 mg L-1 30.00 0.30 ml x 500 mg L-1 75.00 0.30 ml x 750 mg L-1 112.50 1 2 5 8 11 14 15 Brunisol Luvisol-1 Luvisol-2 Anthrosol-1 Anthrosol-2 ∫ H2O ɧ 0.00 0.158 0.11 0.304 0.74 0.619 0.25 0.273 0.26 0.277 MC *Each treatment was performed in six replicates. Three of them were used for sequential extraction analysis by phosphate buffer and the other three were by deionized water. £ Spiked con. = Spiked concentration [mg Cr(VI) kg-1dry soil-1]. ∫ H2O refers to the amount (ml) of distilled water that was added to the 2.00 g of soils for raising soil water to 60% of total porosity. ɧ MC = moisture content [g H2O g-1dry soil-1]. 84 Figure 4.1 Incubation of soil samples in the growth chamber. 4.2.4 Soil Cr analysis The sequential extraction method used for Cr(VI) fractionation analysis in soils entailed multiple phosphate extractions and an alkaline digestion process.209,274 Soils in tubes were shaken with 20 mL of 10 mM K-phosphate buffer extractant (PBE, 5 mM K2HPO4 and 5 mM KH2PO4, pH 7.2) in a reciprocating shaker. After 30 minutes, the soil in each tube was centrifuged for 15 minutes at 8200g at 20 oC (JA-25.50 fixed-angle aluminum rotor, Beckman Coulter Ireland), and each supernatant was filtered through a 0.45 µm membrane filter. Five milliliters of filtered supernatant were taken by pipette for colorimetric Cr(VI) measurement.221 The remaining supernatant was discarded, and the phosphate extraction was repeated up to three times for samples aged from 1 – 30 days and repeated twice for samples aged from 60 – 240 days (Less than 0.6 mg kg-1 Cr(VI) was removed by the third extraction in samples aged over 85 two months). The sum of extracted Cr(VI) from multiple phosphate extractions represented the available Cr(VI) in soils [PBE Cr(VI), mg kg-1].209 An available amount of heavy metal in the soil includes both the quantity in the pore water and the potential available fraction absorbed to the soil matrix.92 After phosphate extractions, the residual soil in tubes was then transferred to 250 mL Erlenmeyer flasks for alkaline digestion to extract the remaining forms of Cr(VI) from soils,220 followed by colorimetric Cr(VI) measurement.221 The results are designated immobile Cr(VI) (mg kg-1). The accuracy of the alkaline extraction method was evaluated in our previous study by analyzing a certified reference sample of soil CRM 041 (LRAB5366).274 Total Cr(VI) (mg kg-1) in soils was calculated as the sum of measured available and immobile Cr(VI). The difference between the added Cr(VI) and total Cr(VI) represents the Cr(III) fraction (mg kg-1) generated from Cr(VI) reduction. Certified reference material of Soil CRM041 (LRAB5366) (Sigma-Aldrich, Canada) was used for accuracy analysis in this study. The measured total Cr(VI) content by the sequential extraction method was 118 ± 1.9 mg kg−1 (112.1 mg kg−1 for the available fraction and 6.0 mg kg−1 for the immobile fraction, respectively), which was within the standard deviation of the certified Cr(VI) content (130 mg kg−1, 13% standard deviation, 91−169 mg kg−1 acceptable interval). The available Cr(VI) determination was also performed by using deionized water at pH 5.7, as above. The sum of Cr(VI) from multiple water extractions represented available Cr(VI) in soils [DWE Cr(VI), mg kg-1]. For the purpose of mass balance analysis, the supernatant from the first water extraction was collected to determine Cr concentrations by ICP-OES (Agilent Technologies 5100), represented as available Cr (mg L-1). The concentration of available Cr(III) 86 (mg L-1) was calculated as the difference between available Cr and available Cr(VI). After distilled water extractions, the residual soil in the tube was digested using the acid (HNO3 & HClO4) block digestion method (SCP Science DigiPREP MS) and the content of Cr was determined by ICP-OES, defined as residue Cr (mg kg-1). The immobilized Cr (mg kg-1) was calculated as the difference between residue Cr and native Cr in soils. 4.2.5 Kinetics modeling A first-order kinetic model, a second-order kinetic model, an Elovich model, a parabolic diffusion model and a fractional power model were applied to describe the kinetics of available Cr(VI) during the aging process (1, 2, 4, 8, 15, 30, 60, 90, 120, 180 and 240 days) in the selected soils (Table 4.2).95 To indicate the proportion of the variation explained by each model, the coefficient of determination (R2) was calculated as R2 = 1 – RSS / CTSS where RSS is the residual sum of squares, CTSS is the corrected total sum of squares.281 Akaike’s information criterion was used to compare and rank multiple competing models, calculated as 𝐴𝐼𝐶𝑐 = 𝑛 [ln( 𝑅𝑆𝑆 2𝑘 (𝑘 + 1) )] + 2𝑘 + 𝑛 𝑛−𝑘−1 𝑛 ( < 40) 𝑘 where n is the samples size, RSS is the residual sum of squares of the model and k is the number of fitted parameters.282 The model having the lowest AICc value represents the best approximating model of the given data.282 87 Table 4.2 Kinetic models used to study available Cr(VI) data in soils with aging.95 Model Equation Parameters ɸ ln 𝑞t = ln 𝑞0 − 𝑘1 t k1, first-order rate constant (d-1) First-order ƞ k2, second-order rate constant (kg mg- Second-order 1/𝑞t = 1/𝑞0 + 𝑘2 t Elovich 𝑞t = (1/𝛽) ln(𝛼𝛽) + (1/𝛽) ln t ∫ 1 d-1) α, initial aging rate (mg kg-1 d-1), and β, aging constant (kg mg-1) kp, diffusion rate constant (kg mg-1d- £ Parabolic diffusion 0.5 𝑞t = 𝑘p t 0.5 + C ), and C, is the constant related to the diffusion layer (mg kg-1) Ʈ Fractional power kF, initial aging rate constant mg kg−1 𝑞t = 𝑘F 𝑡 𝑣 d−v, and v, aging rate coefficient 𝑞0 , available Cr(VI) concentration at t = 0 (mg kg-1); 𝑞𝑡 , available Cr(VI) concentration at time t (mg kg-1); t, the aging time (d). ɸ The first-order model assumes that the rate of a reaction is proportional to available Cr(VI) concentration only.95 In such a case the supply of reducing agents is not limiting. ƞ The second-order model assumes that the reaction rate is proportional to two reactants.95 ∫ The Elovich model has been often useful for describing systems in which the absorbing surfaces is heterogeneous.283 £ The parabolic diffusion model is often used to describe reactions in which diffusion-controlled phenomena are rate-limiting.284 Ʈ The fractional power model is a modified form of the Freundlich equation, and describes a system in which binding strength decreases with the increasing degree of site occupation.285 88 4.2.6 FTIR spectroscopy To identify the functional groups for Cr retention, Fourier Transform Infrared Spectroscopy (FTIR) spectra of soil samples before and after aging were obtained using a Bruker Alpha II spectrometer equipped with an attenuated total reflection accessory. A total of 24 scans were processed for each sample from 4000 to 370 cm−1 with a resolution of 4 cm −1. 4.2.7 Soil respiration measurement For estimating microbial activity the measurement of soil respiration rate (SR) is widely used.286,287,288 Addition of available carbon to soil can cause an increase in the respiration rate (substrate induced respiration, SIR), which reveals the active groups of microorganisms in soil utilizing the substrate.288 In this study, SR and SIR were measured using the sodium hydroxide trapping method.289 The basic experimental unit was moist soil (20 – 40 g of dry weight, wetted to 60% of total porosity) in a 100-mL beaker, 25 ml of 0.25 M NaOH solution in a 50-mL beaker, contained in a 2-liter, wide-mouthed canning jar. The soils for SIR were initially amended with 5 mg glucose g-1 soil,199 and then wetted with deionized water to 60% of total porosity. All the samples were performed in triplicates, with a blank containing only NaOH solution. Jars were sealed and stored in the dark at room temperature for 4 days. After 4 days, 10.00 ml of the NaOH solution was placed into a 50-mL Erlenmeyer flask, added with 1 mL of BaCl2 solution and 2 drops of phenolphthalein indicator. The suspension was titrated with standard 0.25 M HCl until the pink color just disappeared and the suspension appeared milky white. The CO2-C evolved from the soils was calculated as follows: mg CO2-C evolved = (B – V) × (N × E) x 25/10 Divided by the oven-dried mass of the soil (g) used to obtain mg CO2-C/g soil (per unit time) 89 where B = volume (ml) of standard acid required to titrate the trap solution of the control (no soil); V = volume (ml) of standard acid required to titrate the trap solution from samples; N = normality of the acid (mmol/ml); E = equivalent weight of CO2: 6 mg C/mmol NaOH neutralized. SR and SIR are expressed as mg CO2-C kg-1 soil−1 d−1 and specific soil respiration (SSR) is normalized by the sample's organic carbon content into mg CO2-C g SOC−1 d−1. To investigate the influence of spiked Cr(VI) on soil microbial activities, we also determined SR in soils spiked with 150 mg kg-1 Cr(VI) using the sodium hydroxide trapping method as described above. 4.3 Results 4.3.1 Soil properties Selected physiochemical properties of the five soils are summarized in Table 4.3. The Brunisolic and Luvisolic samples are acidic (pH 4.73 – 6.11), while the two Anthrosolic soils from contaminated sites are alkaline (pH 7.17 – 7.92). Particle size analysis showed variability in soil texture. According to the textural triangle,216 samples ranged from sand (Brunisol) through sandy clay loam (Anthrosol-1 and Anthrosol-2) and silty clay loam (Luvisol-1) to clay (Luvisol2). The five soils demonstrated a wide range of total C content, ranging from 0.33% in the Brunisol to 9.37% in the Luvisol-2. Most of the total C in the Luvisol soils was in the organic form (88% in Luvisol-1; 95% in Luvisol-2); while a large proportion of total C in the Anthrosol was in the form of carbonates (43% in Anthrosol-1; 20% in Anthrosol-2) originating from tannery waste.274 Total N content (%) in the soils ranged from 0.02 to 0.50, and CEC (cmol kg-1) in the soils varied from 6.6 to 70.5. Total porosity was highest in Luvisol-2 (72.1%) and lowest in Brunisol (41.1%). 90 Total native Cr concentration in the Brunisolic and Luvisolic samples ranged between 30 and 37 mg kg-1. Anthrosol-1 and Anthrosol-2 collected from the contaminated sites had 145 and 210 mg kg-1 total Cr, respectively. Total Cr(VI) contents in the five soils was < 0.6 mg kg-1, indicating that the majority of Cr in these soils was present as Cr(III). Mean contents of Fe, Al and Mn in soils were variable, with Fe at 1.46 – 2.79 %; Al at 0.98 – 2.49 % and Mn at 209 – 1360 mg kg-1. The concentration of other elements in the soils ranged widely for Ca (0.17 − 0.80%), P (697 – 1759 mg kg-1), S (83 – 498 mg kg-1), Ba (64.5 – 274 mg kg-1) and Pb (<6 – 53.4 mg kg-1). Soluble Cr was low in all the soils (≤ 0.07 mg kg-1). Soluble Ca was higher in Anthrosol-1 (188 mg kg-1) and Anthrosol-2 (37.6 mg kg-1), compared to Brunsiol (0.71 mg kg-1). XRD analysis revealed that the main crystalline components in the soil samples were quite similar, including components of quartz, albite, adularia, muscovite and clinochlore (Figure 4.3). 91 Table 4.3 Selected properties (means with standard deviation, n=3) of five soils used in this study (ND = Not determined). Soil samples Properties Brunisol Luvisol-1 Luvisol-2 Anthrosol-1 Anthrosol-2 pH (H2O) 5.40 (0.02) 5.84 (0.07) 6.11 (0.04) 7.91 (0.09) 7.92 (0.11) pH (CaCl2) 4.73 (0.01) 5.00 (0.05) 5.61 (0.04) 7.27 (0.11) 7.17 (0.03) Total C (%) 0.33 (0.05) 2.18 (0.05) 9.37 (0.33) 0.81 (0.05) 2.03 (0.07) Total N (%) 0.02 (0.01) 0.12 (0.00) 0.50 (0.02) 0.07 (0.00) 0.15 (0.00) ɸ SIC (%) 0.135 0.258 0.427 0.346 0.414 ∫ SOC (%) 0.215 1.92 8.94 0.414 1.616 ƞ CEC 6.6 (0.6) 27.3 (0.8) 70.5 (1.4) 20.9 (0.4) 23.8 (0.2) TP (%) 41.1 57.4 72.3 54.7 55.1 Sand (%) 89.7 (0.5) 19.4 (1.4) 19.9 (1.7) 52.1 (0.3) 50.3 (1.9) Silt (%) 3.1 (0.4) 45.1 (3.8) 38.0 (4.1) 18.4 (2.3) 19.1 (2.2) Clay (%) 7.3 (0.3) 35.5 (2.6) 43.2 (2.3) 29.5 (2.2) 30.5 (1.0) Total elemental analysis Fe (g 100g-1) 1.46 (0.07) 2.77 (0.00) 2.79 (0.07) 2.33 (0.17) 2.49 (0.07) Al (g 100g-1) 0.98 (0.01) 1.67 (0.01) 2.24 (0.05) 2.49 (0.05) 2.43 (0.01) Ca (g 100g-1) 0.17 (0.01) 0.43 (0.00) 0.80 (0.00) 0.36 (0.01) 0.60 (0.00) Mn (mg kg-1) 209 (14) 959 (76) 1360 (13) 759 (16) 1255 (320) P (mg kg-1) 697 (0.5) 1550 (6) 1759 (56) 824 (26) 745 (40) S (mg kg-1) 83 (4.5) 178 (2.0) 498 (1.5) 213 (4.5) 380 (0.5) Cr (mg kg-1) 30 (3.0) 33 (0.5) 37 (0.0) 145 (4.5) 210 (2.5) Cr(VI) < 0.6 < 0.6 < 0.6 < 0.6 < 0.6 Ba (mg kg-1) 64.5 (2.5) 154 (1.0) 274 (5.0) 103 (3.5) 185 (5.5) Pb (mg kg-1) <6 11 (0.0) 11 (0.0) 38.5 (0.5) 53.5 (4.5) Water soluble elements Cr (mg kg-1) < 0.01 < 0.01 < 0.01 0.04 (0.00) 0.07 (0.01) Ca (mg kg-1) 0.71 (0.22) ND ND 188 (21) 376 (19) ɸ Soil inorganic carbon (SIC) (%), which represents the carbonates decomposition (Figure 4.2). Soil organic carbon (SOC) (%) = Total C – SIC. ƞCation exchange capacity (CEC) (cmol kg-1). ∫ 92 Figure 4.2 CaCO3 content in soils as determined by thermogravimetric analysis (TGA). 93 Figure 4.3 XRD powder patterns of the studied samples. Q = quartz SiO2, Ab = albite Na(AlSi3O8), Ad = adularia K(AlSi3O8), M = muscovite, Cl = Clinochlore, Co = Cordierite (Mg2Al4Si5O18). 94 4.3.2 Aging of Cr fractions in soils The changes that occurred in the Cr(III) and Cr(VI) proportions during aging in Brunisol, Anthrosol-1 and Anthrosol-2 are presented in Figure 4.4. The speciation of Cr continually changed during the 240-day incubation period. Available Cr(VI) dominated after one day of incubation, accounting for 66 – 72% of the total added Cr(VI) in Brunisol, 80 – 85% in Anthrosol-1, and 54 – 72% in Anthrosol-2. After 240-d of incubation, 98 – 100% of available Cr(VI) had been transformed in Brunisol, 97 – 100% in Anthrosol-1, and 95 – 100% in Anthrosol-2. In contrast, the size of the immobile Cr(VI) fraction increased with aging time, indicating the transformation of some Cr(VI) from available to immobile forms. After 240-d aging, the proportion of immobile Cr(VI) increased by 22 – 26% of added Cr(VI) in Anthrosol-1 and 26 – 31% in Anthrosol-2. Compared with the Anthrosol soils, much less Cr(VI) was transformed into the immobile fraction (< 7.3%) in the Brunisol soil. As expected, the proportion of Cr(III) rose with aging time. After 240 days, the Cr(III) fraction increased to 90 – 95% of the added Cr(VI) in Brunisol, to 74 – 77% in Anthrosol-1, and to 68 – 72% in Anthrosol-2. Mass balance analysis showed that available Cr(III) was less than 1.6 mg kg-1 in all selected soil samples that were spiked with 75 – 150 mg kg-1 Cr(VI) (Table 4.4). Most of the Cr(III) obtained from Cr(VI) reduction was retained in an immobile form in the soil residues (Table 4.5), consistent with previous studies demonstrating Cr(III) is relatively immobile in soils.93,290 Consequently, Cr(III) mobility in soils is not an issue in this incubation. Cr speciation with aging in the Luvisol is not shown in Figure 4.4 because the spiked Cr(VI) concentration declined to < 1.0 mg kg-1 within 24 hours. 95 Figure 4.4 Changes of Cr fractions as a proportion of total added Cr(VI) during aging in Brunisol spiked with 150, 112.5 and 75 mg kg-1 Cr(VI) [(a) – (c)], in Anthrosol-1 spiked with 75, 30 and 15 mg kg-1 Cr(VI) [(d) – (f)], and in Anthrosol-2 spiked with 112.5, 75 and 30 mg kg-1 [(g) – (i)]. Total added Cr(VI) in soils was fractionated into available Cr(VI), immobile Cr(VI) and Cr(III). 96 Table 4.4 Concentrations (standard deviation) of available Cr, available Cr(VI) and available Cr(III) in selected samples (n = 3). Treatments Brunisol + 75 Cr(VI) Brunisol + 112.5 Cr(VI) Brunisol + 150 Cr(VI) Anthrosol-1 + 75Cr(VI) Anthrosol-2 + 75Cr(VI) Anthrosol-2 + 112.5Cr(VI) Aging (d) Available Cr (mg L-1) Available Cr(VI) (mg L-1) Available Cr(III) (mg L-1) Available Cr(III) (mg kg-1) 60 0.47(0.00) 0.41 (0.00) 0.06 0.6 120 0.06(0.00) 0.14 (0.14) -0.08 -0.8 180 0.04(0.00) 0.01(0.00) 0.02 0.2 240 0.01(0.00) 0.00(0.00) 0.01 0.1 60 1.26(0.01) 1.12(0.20) 0.13 1.3 120 0.21(0.01) 0.20 (0.00) 0.01 0.1 180 0.20(0.01) 0.13(0.01) 0.07 0.7 240 0.13(0.00) 0.02(0.00) 0.11 1.1 60 2.39(0.80) 2.23(0.00) 0.16 1.6 120 0.77(0.02) 0.79(0.03) -0.01 -0.1 180 0.46(0.00) 0.41(0.00) 0.05 0.5 240 0.28(0.00) 0.22(0.00) 0.06 0.6 120 0.61(0.00) 0.63(0.00) -0.02 -0.1 180 0.27(0.00) 0.27(0.00) 0.00 0.5 240 0.22(0.01) 0.20(0.00) 0.02 0.6 120 0.37(0.00) 0.40(0.00) -0.03 -0.3 180 0.31(0.00) 0.36(0.00) -0.05 -0.5 240 0.29(0.00) 0.27(0.00) 0.02 0.2 120 0.61(0.57) 0.57(0.00) 0.04 0.4 180 0.57(0.55) 0.55(0.00) 0.02 0.2 240 0.49(0.01) 0.50(0.00) -0.01 -0.1 97 Table 4.5 Mass balance analysis of immobile Cr concentrations in mg kg-1 (standard deviation) in selected soil samples (n = 3) Treatments Brunisol + 75Cr(VI) Brunisol + 112.5Cr(VI) Brunisol + 150Cr(VI) Anthrosol-1 + 75Cr(VI) Anthrosol-2 + 75Cr(VI) Anthrosol-2 + 112.5Cr(VI) Aging (d) Residue 120 98.4(4.5) 68.4 180 103 (1) 240 ɸ Recovery (%) Im Cr(VI) 74.9 91.3 4.6 (1.9) 63.8 73.2 74.8 97.8 5.5 (0.0) 67.7 108 (8) 78.1 75.0 104.2 3.4 (1.1) 74.8 120 125 (16) 95.2 112.2 84.8 9.0 (0.9) 103.2 180 134 (10) 104 112.2 92.4 7.7 (0.3) 104.5 240 130 (7) 99.8 112.5 88.7 7.2 (1.1) 105.3 120 152 (3) 122 146.8 83.3 14.4 (0.2) 132 180 157 (23) 127 149.5 84.7 14.1 (0.3) 135 240 159 (3) 129 149.9 85.7 10.9 (0.9) 139 120 213 (2) 68.3 71.8 85.2 20.1 (3.6) 48.2 180 217 (8) 71.8 74.5 86.4 17.0 (0.6) 54.9 240 222 (4) 76.5 74.9 102.2 16.7 (2.7) 60.0 120 286 (12) 75.5 71.0 106.4 19.3 (2.3) 56.2 180 279 (10) 68.6 71.4 86.1 18.7 (1.3) 49.9 240 286 (6) 76.4 71.9 106.2 19.1 (1.7) 57.2 120 308 (6) 97.9 106.8 91.6 30.5 (2.3) 67.3 180 308 (19) 98.0 107.0 91.5 30.9 (3.9) 67.1 240 311 (13) 101 106.8 94.6 30.4 (1.7) 70.7 Cr ƞ Im Cr £ Cal im Cr ƞ Ʈ Im Cr(III) Immobile Cr (mg kg-1) = residue Cr – native Cr; Native Cr (mg kg-1) was 30 in Brunisol, 145 in Anthrosol-1 and 210 in Anthrosol-2 (Figure 4.3). £ Calculated immobile Cr (mg kg-1) = total added Cr(VI) – soluble Cr; ɸ Recovery (%) = immobile Cr /calculated immobile Cr *100; Ʈ Im Cr(III) (mg kg-1) = Im Cr – Im Cr(VI). 98 4.3.3 Kinetics of available Cr(VI) in soils during aging The dynamics of available Cr(VI) were essentially identical regardless of initial spiking concentrations for Brunisol, Anthrosol-1 and Anthrosol-2 (Figure 4.5). The amount of available Cr(VI) sharply declined in the initial 30 days and subsequently declined slowly (in Brunisol and Anthrosol-1) or remained almost constant (in Anthrosol-2) with subsequent aging. Exogenous Cr(VI) aging in the three soils was modeled by first-order, second-order, Elovich, parabolic diffusion, and fractional power models (Figure 4.6). The values of the rate constant, correlation coefficients, and AICc are presented in Table 4.6. The Elovich model received the lowest AICc score representing the most parsimonious approximation of the data for Brunisol and Anthrosol1. In comparing reaction rates, the constants α and β can be used, and a decrease in the value of β and/or an increase in the value of α represent an increased reaction rate.283 The value of β decreases with increasing spiking concentrations in Brunisol and Anthrosol-1, indicating the rate of transformation for available Cr(VI) increases with rising spiked Cr(VI) concentrations. With 75 mg kg-1 Cr(VI) added, the reaction rate in Anthrosol-1 (β = 0.068 kg mg-1) was faster than in Brunisol (β = 0.118 kg mg-1). The fractional power model (R2 = 0.993; AICc = 45.2) was selected as the best fit model for Anthrosol-2 at 112.5 mg kg-1 Cr(VI), whereas the first-order kinetic model (R2 = 1.000; AICc = 36.7 – 45.4) fitted best for Anthrosol-2 at 75 and 30 mg kg-1 Cr(VI). 99 Figure 4.5 Changes of available Cr(VI) content (mg kg-1) during the whole aging period in (a) Brunisol spiked with 150, 112.5 and 75 mg kg-1 Cr(VI); (b) Anthrosol-1 spiked with 75, 30 and 15 mg kg-1 Cr(VI); (c) Anthrosol-2 spiked with 112.5, 75 and 30 mg kg-1 Cr(VI). PBE = phosphate buffer extraction; DWE = deionized water extraction. 100 Figure 4.6 Cr(VI) with aging and the fit of data by kinetic models in Brunisol spiked with 150, 112.5 and 75 mg kg-1 Cr(VI) [(a) – (c)], in Anthrosol-1 spiked with 75, 30 and 15 mg kg-1 Cr(VI) [(d) – (f)], and in Anthrosol-2 spiked with 112.5, 75 and 30 mg kg-1 [(g) – (i)]. 101 Table 4.6 The fitted coefficient (R2) and Akaike’s information criterion value (AICc) and parameters of kinetic models on available Cr(VI) aging process. Models First-order Secondorder Elovich Parabolic diffusion Fractional power Brunisol Anthrosol-1 Anthrosol-2 𝑞0 150 112.5 75 75 30 15 112.5 75 30 R2 0.846 <0 0.908 0.958 0.932 <0 0.861 1 1 AICc 57.7 84.6 45.9 75.7 137 146 75.9 45.4 36.7 k1 0.121 0.182 0.220 0.034 0.022 0.175 0.150 0.265 0.413 R2 0.952 0.715 0.964 0.925 0.43 0.283 0.976 0.995 0.882 AICc 46.1 64.2 36.4 82.1 160 135 56.3 164 158 k2 0.0012 0.0009 0.0049 0.0009 0.011 0.044 0.0026 0.012 0.122 R2 0.995 0.993 0.936 0.985 1 0.996 0.943 1 0.666 AICc 25.5 29.8 45.3 66.9 79.1 80 68.8 98 172 α 5048 3014 1686 4020 429 162 1973 520 30 β 0.054 0.073 0.118 0.068 0.56 1.48 0.092 0.346 2.88 R2 0.958 0.86 0.666 0.96 1 0.993 0.832 1 0.419 AICc 47.8 60.2 61.9 77.8 83.6 85.7 80.7 107 178 kP -5.60 -3.24 -2.23 -3.56 -0.25 -0.09 -3.60 -0.85 -0.09 C 84.9 44.5 31.4 56.9 3.84 1.45 44.2 11.4 0.88 R2 0.912 0.795 0.885 0.738 0.419 0.262 0.993 1 0.998 AICc 55.2 64 51.2 98.6 164 137 45.2 109 113 kF -0.46 -0.46 -0.71 -0.67 -1.2 -1.3 -0.53 -2.0 -2.8 v 124 103 56.7 181 291 92.5 101 538 336 𝑞0 , available Cr(VI) concentration at t = 0 (mg kg-1). k1, first-order rate constant (d-1); k2, second-order rate constant (kg mg-1 d-1); α, initial aging rate (mg kg-1 d-1), and β, aging rate constant (kg mg-1) of the Elovich model; kp, rate constant (kg mg-1d-0.5), and C, is the constant (mg kg-1) of the parabolic diffusion model; kF, initial aging rate constant (kg mg-1d-0.5), and v, aging rate coefficient of fractional power model. R2, the coefficient of determination; AICc, Akaike’s information criterion score. 102 The two Luvisols had high initial Cr(VI) reduction rates. Consequently, to determine the progression of rates over a greater cumulative concentration range, an additional experiment was employed for Luvisol-1 and Luvisol-2. This experiment involved spiking 0.3 mL of 1000 mg Cr(VI) L-1 K2Cr2O7 solution to air-dried soils (2.00 g oven dry equivalent) and monitoring the change of available Cr(VI) over time. When there was no detectable Cr(VI) remaining (< 0.6 mg kg-1), another aliquot of Cr(VI) was added (Figure 4.7). Limited (< 2 mg kg-1) immobile Cr(VI) was detected in this experiment, suggesting the reduction of Cr(VI) to Cr(III) may be so fast as to prevent ancillary immobilization reactions. The amount of Cr(VI) (mg kg-1) reduced by a particular soil is defined as the capacity of the soil for the reduction of soluble Cr(VI) 264. As expected, Luvisol-2 containing 7% more total carbon has greater reduction capacity than Luvisol-1 (≥1500 mg kg for -1 Luvisol-2; ≥450 mg kg-1 for Luvisol-1) (Figure 4.7a). Plots of the log of available Cr(VI) versus time are linear, suggesting Cr(VI) reduction in Luvisol follows the first-order reaction kinetics (0.975 < R2 < 0.996), except for the 10th spiking of Cr(VI) in Luvisol-2 where the second order kinetics became apparent (R2 = 0.956) (Figure 4.7b; Table 4.7). Relative to the concentration of added Cr(VI), the concentrations of electron donors in the Luvisols were sufficiently high and effectively constant during the reduction. The reduction rate decreased with each successive addition of Cr(VI). Therefore, Cr(VI) reduction in Luvisol soils could be described as pseudo-first-order kinetics. Reduction rate in Luvisol-2 was relatively fast (k1 = 4.08 d-1), more than two times higher than in Luvisol-1 (k1 = 1.52 d-1). 103 Figure 4.7 (a) Changes of available Cr(VI) content (mg kg-1) and (b) first-order plots of ln[available Cr(VI)] in Luvisol-1 and Luvisol-2 with multiple Cr(VI) spiking as a function of time. [Cr(VI)]0 = 150 mg kg-1 for each spike; The slope of the plots is equal to the first-order rate constant. Vertical scales are different. Table 4.7 Fitted linear regressions of first-order and second-order reactions for Luvisol-1 and Luvisol-2 with multiple spiking of 150 mg kg-1 Cr(VI). Soil First-order Second-order R2 k1 R2 k2 Luvisol-1+150*1st 0.993 1.52 0.661 3.43 Luvisol-1+150*2nd 0.996 0.429 0.660 0.125 Luvisol-1+150*3rd 0.980 0.224 0.799 0.028 Luvisol-2+150*1st 0.988 4.08 0.921 0.372 Luvisol-2+150*2nd 0.991 1.98 0.847 0.198 Luvisol-2+150*3rd 0.984 1.59 0.755 0.177 Luvisol-2+150*4th 0.991 1.05 0.711 0.389 Luvisol-2+150*8th 0.994 0.454 0.829 0.084 Luvisol-2+150*9th 0.975 0.344 0.608 0.183 Luvisol-2+150*10th 0.849 0.146 0.956 0.011 k1, aging rate constant of the first-order kinetic model (d-1); k2, aging rate constant of the second-order kinetic model (kg mg-1d-1); 1st – 10th refers to the spiking times for Luvisol-1. 104 4.3.4 Changes of immobile Cr(VI) and Cr(III) fractions in soils during aging The changes of immobile Cr(VI) and immobile Cr(III) fractions during aging of the three soils are shown in Figure 4.8. The concentration of immobile Cr(VI) increased in the initial 30 days and remained constant for Brunisol and Anthrosol-2 but decreased slightly for Anthrosol-1 during the remainder of the aging experiment, indicating the slow reduction of immobile Cr(VI) in Anthrosol-1. The kinetics of Cr(VI) immobilization in soils was nonlinear and did not appear to conform to any of the tested kinetic models. The Cr(III) fraction increased rapidly in the first 30 days but increased slowly after that. The amount of Cr(III) formed in the Brunisol was significantly higher than the amount of immobile Cr(VI) during the entire aging process, suggesting reduction was the major reaction. Unlike the Brunisol, Cr(VI) immobilization in the Anthrosol-1 was faster than reduction in the first 15 days, but afterwards, reduction dominated. Figure 4.8 Changes of immobile Cr(VI) and immobile Cr(III) contents (mg kg-1) during the whole aging period in (a-c) Brunisol spiked with 150, 112.5 and 75 mg kg-1 Cr(VI); (d-f) Anthrosol-1 spiked with 75, 30 and 15 mg kg-1 Cr(VI); (g-i) Anthrosol-2 spiked with 112.5, 75 and 30 mg kg-1 Cr(VI). 105 4.3.5 FTIR results The broad band around 3200 – 3400 cm-1 is attributed to hydroxyl (O-H) groups.291 The weak peak around 1633 cm-1 could be assigned to C=O stretching of amide groups, while the band around 1420 cm-1 may originate from the symmetric stretching vibrations of COO− anions.292 The strong band around 1005 cm-1 could be assigned to either Si–O–Si or Si–O–C structures.293 As shown in Figure 4.9, band shifts (from 3328 to 3352 cm-1 and from 1633 to 1622 cm-1) were observed in Luvisol-2 after spiking with 1500 mg kg-1 Cr(VI), suggesting the hydroxyl and amide groups were involved in Cr retention. The band shift from 1001 to 1005 cm-1 in Luvisol-1 indicates that Cr likely is also bound to clay minerals. In Anthrosol-2, COO− anions could be possible for Cr complexation. Figure 4.9 FTIR spectra of Anthrosol-2 spiked with 112.5 mg kg-1, Luvisol-1 spiked with 450 mg kg-1 and Luvisol-2 spiked with 1500 mg kg-1 after aging. 106 4.4 Discussion 4.4.1 Transformations of Cr(VI) in soils with aging It has been recognized that soil contaminants, whether organic (e.g. petroleum hydrocarbons) or inorganic (e.g. metal ions), are usually less than 100% bioavailable.294,295 After soluble metal or metalloids enter soils, many reactions occur such as adsorption, micropore diffusion, precipitation, and crystallization,296 with redox reactions being especially important for elements such as Cr, Se, and As.297 The operating mechanisms may change with reaction time, and with aging, resulting in differences in the availability of a contaminant in soils.296 Results in this study are consistent with this consensus respecting the aging of Cr in five soils. The concentration of available Cr(VI) continually declined in all tested soils with increasing time, although transformation kinetics varied among the soils (Figure 4.5; Figure 4.7). The decrease of available Cr(VI) in the Luvisol soils was relatively rapid, following the pseudo first-order kinetics (Figure 4.7). A fast decline of available Cr(VI) in Brunisol, Anthrosol-1 and Anthrosol-2 was recorded in the initial period (30 days) followed by a slow decrease thereafter with aging (Figure 4.5), which is consistent with previous reports on selenium298, arsenic299 and chromium300. Li et al. (2016) reported that soil available selenium decreased rapidly in the first 42 days and at a much slower rate with subsequent time.298 Wang et al. (2017) reported a sharp decrease of available arsenic in soils with aging for 30 days and a slight decrease for the remainder of one-year incubation.299 Yang et al. (2019) reported that extracted Cr(VI) in soils declined rapidly in the initial 35 days and then slowly reached an equilibrium.300 This biphasic change was inadequately described as a single first-order kinetic reaction (Table 4.6), but often is interpreted as a combination of two or three simultaneous first-order reactions.283 In the present study, the decrease of available Cr(VI) 107 was best fit by the Elovich model in Brunisol and Anthrosol-1, and by fractional power model in Anthrosol-2, respectively (Figure 4.6; Table 4.6). Sequential extraction analysis further demonstrated that the decline of available Cr(VI) with aging was attributed to its reduction to Cr(III) and/or its immobilization (Figure 4.4; Figure 4.8), with the distribution between these two fates varying among soils. The reduction of Cr(VI) to Cr(III) is prevalent in soils, partly because of the high pE (8.2 at pH 7) associated with Cr(VI)/Cr(III) transformation.71 Various components present in soils can serve as electron donors to reduce Cr(VI), including organic matter (soluble and insoluble), Fe(II) ions and sulfides.10 1 − Fendorf et al. (2000) represented the reduction of Cr(VI) by ferrous iron as: Fe2+ (aq) + 3 HCrO4 + 4 5 H2 O ↔ 3 Cr0.25 Fe0.75 (OH)3 + 3 H + , ∆Go (pH 7) = −147 kJmol−1.13 Lan et al. (2005) have − + shown the stoichiometry for Cr(VI) reduction by sulfide is: CrO2− 4 + 3HS + 7H ↔ 2Cr(OH)3 + 3S(s) + 2H2 O.301 Soil organic matter functional groups have been reported to participate in Cr(VI) reduction in the order of carbonyl (C=O) > phenol (C-O) > hydroxyl (O-H) > methyl (C-H).73,302 The functional group of phenol has been demonstrated to play crucial roles for scavenging Cr(VI).75,302 Cr(VI) reduction by dihydroxyphenol can be represented by the + 3+ following ionic reaction: Cr2 O2− + 3C6 H4 O2 + 7H2 O.303 In 7 + 8H + 3C6 H4 (OH)2 ↔ 2Cr this study, FTIR results showed that hydroxyl groups and amide groups were likely involved in binding with Cr in Luvisol-2 and carboxyl groups reduced Cr(VI) in Anthrosol-2 (Figure 4.9). Microbial Cr(VI) reduction in soils could be catalyzed directly by enzymatic activities or indirectly by the metabolic end-products of iron minerals such as Fe(II).247,304 In the present study, soil microbial activity, as indicated by soil respiration rate (SR), was highly variable (Table 4.8). Spiking with Cr(VI) significantly decreased microbial activities in all of the soils 108 (Table 4.8), which was in agreement with a previous report.305 Microbial Cr(VI) reduction was not expected to occur in Brunisol and Anthrosol-1 since the microbial activity decreased to less than 12 mg CO2-C kg-1d-1 4 days after amendment with 150 mg kg-1 Cr(VI) (Table 4.8). Substrate-induced soil respiration results indicated the microbial community was active in Luvisol-2 or potentially active in Luvisol-1 and Anthrosol-2 (Table 4.8), suggesting a possibility of microbial Cr(VI) reduction. Cr(VI) reduction was observed in sediments incubated with Cr(VI) for 30 days, with phylogenetic composition and structure of microbial communities changed and involved in Cr(VI) reduction.306 Table 4.8 Mean (standard deviation) of soil respiration rates in the soils (n = 3). Soil samples Substrate induced Respiration rate (CO2-C mg kgrespiration rate (CO21 soil-1d-1) C mg kg-1soil-1d-1) Specific respiration rate (CO2-C mg g1 organic C-1d-1) No Cr(VI) Add Cr(VI) No Cr(VI) No Cr(VI) Brunisol 28.2 (3.8) [a] 4.5 (0.0) [b] 36.3 (1.5) [a] 13.1 (1.8) Luvisol-1 85.8 (2.5) [a] 29.5 (3.3) [b] 120 (3) [c] 4.5 (0.1) Luvisol-2 286 (5) [a] 248 (6) [b] 297 (5) [a] 3.2 (0.0) Anthrosol-1 37.4 (6.6) [a] 11.3 (2.3) [b] 51.3 (3.2) [a] 9.0 (1.6) Anthrosol-2 59.2 (2.4)[a] 35.2 (0.0) [b] 78.3 (2.2) [c] 3.7 (0.1) Note: treatments within a soil not sharing the same letters are significantly different (p < 0.05) by ANOVA. Despite the prevalence of reduction, however, some Cr(VI) was immobilized without reduction. The highest proportions of immobile Cr(VI) were formed in Anthrosol-1 (22 – 26% of total spiked) and Anthrosol-2 (26 – 31% of total spiked) with aging (Figure 4.4). The highest concentrations of soluble Ca (188 – 376 mg kg-1) were in these soils (Table 4.3) suggesting the immobilization of Cr(VI) in the Anthrosols was via precipitation of moderately soluble CaCrO4 (Ksp 7.1 × 10−4). This suggestion is consistent with reports that moderately-to-sparingly soluble 109 precipitates of Cr(VI) were formed with aging in the presence of cations such as Ba2+, Pb2+ and Ca2+.215,274 Compared to the Anthrosols, much less Cr(VI) was redistributed to the immobile fraction (< 7.3%) in the Brunisol (Figure 4.4). The immobilization mechanism in the Brunisol could be coprecipitation of Cr(VI) on the surface of Fe oxides.11,307 Crystallization of Cr(VI) with Fe oxides was not expected to occur in the 240-day duration of this incubation but may account for the increasing resistance to release of Cr(VI) with aging in soils.214 Immobile Cr(VI) formation was most prevalent in soils with slow reduction processes (Figure 4.4). Although the amount of Cr(III) generated differed among soils and treatments, in all cases essentially all Cr(III) was immobilized with available Cr(III) < 2 mg kg-1 by day 240 (Table 4.4 – 4.5). These data indicate Cr(VI) reduction and immobilization were occurring concurrently, and in some cases competing for available Cr species. 4.4.2 Dominant impact factors influencing Cr(VI) transformations with aging The incubation condition was the same but contrasting kinetics of Cr(VI) transformations were observed among soils subjected to aging (Figure 4.5; Figure 4.7). As shown in Table 4.3, pH, TOC, particle size distribution, and cation exchange capacity differed considerably among the soils. Thus, the variations in Cr(VI) transformations during aging may be ascribed to variation in the properties of the soils that were used. The reduction rate of available Cr(VI) to Cr(III) was fast in acidic Luvisol-1 (TOC = 8.9%) and Luvisol-2 (TOC = 1.9%), but much slower in Brunisol with low TOC at 0.2% and in Anthrosols with high pH at 7.9 (Figure 4.5; Figure 4.7). These results indicated available Cr(VI) reduction in soils was dependent mostly on total organic carbon content and on pH, which is consistent with previous reports.71,264,300,308 Kožuh et al. (2000) reported Cr(VI) reduction was 110 much faster in peat soils (TOC = 41.1%) than in Cambisols (TOC = 0.22%).264 Wittbrodt and Palmer (1997) reported the reduction by soil humic acid was on the order of days at pH 2 but on the order of months at pH 7.12 In contrast, when spiked with 75 mg Cr(VI) kg-1, 3.4 mg kg-1 of immobile Cr(VI) was formed in Brunisol and 19.1 mg kg-1 of immobile Cr(VI) was formed in Anthrosol-2, but almost no Cr(VI) (< 2.0 mg kg-1) was detected in the Luvisols spiked with 150 mg Cr(VI) kg-1 (Figure 4.5; Figure 4.7). The above results suggested that Cr(VI) is expected to be immobilized in soils when Cr(VI) reduction is slow. The amount of immobilized Cr(VI) varied among the soils with greater immobilization in soils with higher pH and higher concentrations of soluble cations (such as Ca2+). Leaching of Cr(VI) in soils was significantly lower in the presence of CaCl2 compared to the absence of CaCl2, due to the formation of CaCrO4 at higher soil pH.309 Therefore, we can conclude that soil properties influence Cr(VI) transformations in two aspects, with soil pH and soil organic carbon facilitating Cr(VI) reduction, whereas the concentration of soluble cations (such as Pb2+, Ba2+, and Ca2+) and Fe oxides enhances Cr(VI) immobilization. 4.4.3 Multi-reaction modeling To explore concurrent aging processes, a multi-reaction model was developed to describe flow rates of Cr among Cr fractions with aging, implemented in the MATLAB Simulink modeling environment. MATLAB Similink toolbox has been applied to solve the governing differential equations of many dynamic systems.310,311 The main conditions for modeling simplification were as follows (Figure 4.10): 111 Figure 4.10 Diagram of the multi-reaction model on Cr(VI) transformations in soils. (i) The adsorption of Cr(VI) in soils was reversible.272,14 A bench technique was used to study the adsorption isotherms according to Choppala et al. (2018).312 Indicated by correlation coefficient, Freundlich equation (R2 = 0.993 – 0.999) fitted better than Langmuir equation (R2 = 0.979 – 0.997) in the three tested samples (Figure 4.11; Table 4.9), which is consistent with previous studies.312 Freundlich model received the lowest AICc score (Table 4.9), indicating that this model is the most parsimonious model for the sorption isotherm data. Therefore, the reaction between soluble and adsorbed Cr(VI) was described by Freundlich equation in the multi-reaction model, written as equation (1). By definition, the sum of soluble and adsorbed Cr(VI) represented available Cr(VI) in soils 92 is written as equation (2). 112 Figure 4.11 Experimental data and sorption isotherms of Cr(VI) onto (a) Brunisol, (b) Anthrosol-1, and (c) Anthrosol-2 (2.00 g air-dried soils in 20 mL K2Cr2O7 solution in the range of 2 – 100 mg Cr(VI)/L ; contact time 16 h; room temperature 24 °C): Linear model (red solid lines), Freundlich model (green solid lines), and Langmuir (blue solid lines). 113 Table 4.9 Isotherm constants, correlation coefficients (R2) and Akaike’s information criterion (AICc) obtained for models for Cr(VI) sorption onto soils. Linear Freundlich Langmuir Brunisol Anthrosol-1 Anthrosol-2 Kd 0.922 0.713 0.687 R2 0.910 0.992 0.989 AICc 58.7 22.0 29.2 KF 2.347 0.874 0.880 1/n 0.714 0.922 0.927 R2 0.993 0.994 0.999 AICc 34.8 23.1 12.8 b 0.0219 0.0029 0.0028 Xm 71.77 261.9 274.5 R2 0.975 0.986 0.997 AICc 45.0 28.3 21.5 (ii) Available Cr(VI) was being immobilized with aging (Figure 4.4). We described reactions between available and immobile Cr(VI) fractions using first-order kinetics and set the immobilization rate for soluble Cr(VI) equal to that for adsorbed Cr(VI) (k3 = k4). (iii) All of the Cr(VI) fractions were being reduced with aging, and reactions were described using first-order kinetics.264 The reduction rate of available Cr(VI) was set to be two times faster than that of immobile Cr(VI) (k5 = k6 = 2k7). The oxidation of Cr(III) to Cr(VI) in soils was considered to be insignificant under these conditions and was not included in the model.264 Mass balance equations for the above reactions used for modelling were written as equation (3) – (6). 114 𝐶𝑟 6+ (𝑎𝑑) = 𝑘𝐹 ∗ 𝐶𝑟 6+ (𝑙)1/𝑛 (1) 𝐶𝑟 6+ (𝑙) + 𝐶𝑟 6+ (𝑎𝑑) = 𝐶𝑟 6+ (𝑎𝑣) (2) 𝑑𝐶𝑟6+ (𝑙) 𝑑𝑡 = − 𝑘1 𝐶𝑟 6+ (𝑙) + 𝑘2 𝐶𝑟 6+ (𝑎𝑑) − 𝑘3 𝐶𝑟 6+ (𝑙) − 𝑘5 𝐶𝑟 6+ (𝑙) 𝑑𝐶𝑟6+ (𝑎𝑑) 𝑑𝑡 𝑑𝐶𝑟6+ (𝑖𝑚) 𝑑𝑡 𝑑𝐶𝑟3+ (𝑖𝑚) 𝑑𝑡 (3) = 𝑘1 𝐶𝑟 6+ (𝑙) − 𝑘2 𝐶𝑟 6+ (𝑎𝑑) − 𝑘4 𝐶𝑟 6+ (𝑎𝑑) − 𝑘6 𝐶𝑟 6+ (𝑎𝑑) (4) = 𝑘3 𝐶𝑟 6+ (𝑙) + 𝑘4 𝐶𝑟 6+ (𝑎𝑑) − 𝑘7 𝐶𝑟 6+ (𝑖𝑚) (5) = 𝑘5 𝐶𝑟 6+ (𝑙) + 𝑘6 𝐶𝑟 6+ (𝑎𝑑) + 𝑘7 𝐶𝑟 6+ (𝑖𝑚) (6) 𝑘3 = 𝑘4 (7) 𝑘5 = 𝑘6 = 2𝑘7 (8) where Cr6+(l), Cr6+(ad), Cr6+(av), Cr6+(im) and Cr3+(im) are the concentrations of soluble Cr(VI), adsorbed Cr(VI), available Cr(VI), immobile Cr(VI) and immobile Cr(III) fractions in soils (mg kg-1); k1 (d-1) and k2 (d-1) are the forward and reverse rate coefficients of adsorption; k3 (d-1) and k4 (d-1) are immobilization rate coefficients of soluble and adsorbed Cr(VI), respectively; k5 (d-1), k6 (d-1) and k7 (d-1) are reduction rate coefficients of soluble, adsorbed and immobile Cr(VI), respectively. Input data for the model included Cr6+(av), Cr6+(im), Cr3+(im), kF and n. The change of reaction flow rate and rate coefficients with aging were derived by solving the equations simultaneously in the multi-reaction model. The Simulink procedure scheme for the presented model was shown in Figure 4.12. 115 Figure 4.12 Simulink procedure scheme for the presented model. Cm, measured available Cr(VI) (mg kg-1); C1, C2, C3, and C4 are the concentrations of soluble Cr(VI), adsorbed Cr(VI), immobile Cr(VI) and immobile Cr(III) fractions in soils (mg kg-1); k1 (d-1) and k2 (d-1) are the forward and reverse rate coefficients of adsorption; k3 (d-1) and k4 (d-1) are immobilization rate coefficients of soluble and adsorbed Cr(VI), respectively; k5 (d-1), k6 (d-1) and k7 (d-1) are reduction rate coefficients of soluble, adsorbed and immobile Cr(VI), respectively. 116 As observed in Figure 4.13(a)-(c), the concentrations of available Cr(VI), immobile Cr(VI) and immobile Cr(III) fractions with aging in the model by the interpolation presented very good conformity with experimental results. Figure 4.13(d)-(f) exhibited the reaction flow rates among various Cr fractions with aging. In Anthrosol-1, two reactions were dominant in the initial 15 days including soluble Cr(VI) immobilization and soluble Cr(VI) reduction, with the immobilization flow rate (k3C1) being the fastest; after 15 days, three reactions dominated: soluble Cr(VI) reduction, immobile Cr(VI) reduction and soluble Cr(VI) immobilization, with the reduction flow rate (k5C1) being the fastest. In Brunisol and Anthrosol-2, the reduction flow rate (k3C1) was the fastest during the entire aging process. A sharp decline of the immobilization flow rate (k3C1 < 0) was observed in Brunisol at 8th – 15th days of incubation, which may reveal a fast dissolution of immobile Cr(VI) in Brunisol. All of the reaction rate coefficients (k3 – k7) change with time as shown in Figure 4.13(g)-(i), indicating the transformations among Cr fractions in soils were not controlled by one process but by multiple aging processes. These multiple processes are chemical or microbial reduction, precipitation and crystallization as discussed above. Consequently, the presented model has the ability to describe the dynamics of concurrent Cr(VI) aging reactions in natural soils. The model can be improved if mechanisms and kinetics of various Cr(VI) immobilization reactions in soils with aging can determined and incorporated. 117 Figure 4.13 The change of Cr concentrations (mg kg-1) in various fractions [Cr6+(l), Cr6+(ad), Cr6+(av), Cr6+(im) and Cr3+(im)] [(a) - (c)], reaction flow rates (mg kg-1d-1) [(d) - (f)] and rate coefficients of first-order kinetic model [(g) - (f)] with aging in Anthrosol-1 spiked with 75 Cr(VI) mg kg-1 and Anthrosol-2 spiked with 112.5 Cr(VI) mg kg-1. k3C1: Cr6+(l)→Cr6+(im); k4C2: Cr6+(ad)→Cr6+(im); k5C1: Cr6+(l)→Cr3+(im); k6C2: Cr6+(ad)→Cr3+(im); k7C3: Cr6+(im)→Cr3+(im). 118 4.5 Summary Aging is an important factor controlling metal availability and toxicity in soils. Following spiking soils with Cr(VI), the available Cr(VI) in soils declined sharply in the initial days followed by a slow decrease over the remainder of the 240-day aging period. The decline was associated with an increase in immobile Cr(VI) and Cr(III) fractions. We conclude that aging in these soils was associated with concurrent Cr(VI) reduction and immobilization, which were competing for Cr(VI) species. The fastest rate of decrease in available Cr(VI) was observed in Luvisol-2 with no immobile Cr(VI) formed, whereas the slowest rate of decrease in available Cr(VI) was found in Anthrosol-1 with large amount of immobile Cr(VI) formed, indicating the importance of soils properties in Cr(VI) transformations in natural soils. Using a multi-reaction model, the dynamic changes of various Cr species in soils with aging were well described. 119 Chapter 5 METAGENOMICS REVEALS MICROBIAL COMMUNITY IN CR-CONTAMIANTED SOILS AND ITS DIRECT USE FOR MICROBAIL CR(VI) REDUCTION Abstract In this study, we tested the hypothesis that Cr-contaminated soils hosted microorganisms capable of converting toxic Cr(VI) to non-toxic Cr(III). Four soil samples were utilized, having Cr contents from 296 mg kg-1 in S1-6 to 3141 mg kg-1 in S3-2. Shotgun metagenomic sequencing results showed the accumulation of high levels of Cr in S3-2 led to the increased abundance of Cr resistant and reducing microorganisms: Proteobacteria (69.9%) at phylum level, Betaproteobacteria (39.1%) at class level, Massilia (12.6%) and Bacillus (0.57%) at genus level, compared with S1-6. In a batch experiment Cr(VI) concentrations of 10 or 20 mg L-1 were completely reduced to Cr(III) following addition of 1.0 g of Cr-contaminated soils to 20 mL of K2Cr2O7 solution anaerobically or aerobically under the condition of: pH 7.8 – 8.0, 30 oC, and 0.2 g L-1 of Na-acetate. The amount of Cr(VI) removed was highest (29.0 mg L-1) by S3-2 microbial consortia at 40 mg Cr(VI) L-1 and was highest (15.8 mg L-1) by S1-6 microbial consortia at 20 mg Cr(VI) L-1, which is consistent with microbial community analysis. These results demonstrate that microbial consortia in high Cr-contaminated soils could be directly used for bioremediation of Cr(VI)-polluted environments. Keywords: high Cr-contaminated soils; microbial consortia; microbial Cr(VI) reduction; in vitro 120 Graphical Abstract 121 5.1 Introduction Chromium contamination in soils has been a worldwide environmental problem due to the extensive uses of Cr chemicals in industries such as metal plating, leather tanning, wood preservation and pigment.6 Chromium occurs in the environment in two oxidation states, Cr(III) and Cr(VI), which have contrasting toxicity and mobility. Cr(VI) is known to be carcinogenic and highly mobile, whereas Cr(III) is nontoxic and relatively immobile in soils.10 Metals do not degrade or decay in soils; therefore converting the toxic Cr(VI) to nontoxic Cr(III) has been the key strategy for remediating of Cr contaminated sites to meet regulatory standards and protect human health.6 Microbial reduction of Cr(VI) to Cr(III) is a potential technique in remediation of Crcontaminated sites which would be cost-effective and eco-friendly. Since the first case of microbial reduction of Cr(VI) by Pseudomonas strain was reported by Romanenko & Koren’Kov (1977) that used Cr(VI) as a terminal electron accepter,156 numerous Cr(VI) reducing microbial strains have been isolated from various environments that have shown excellent capacity to reduce Cr(VI) to Cr(III). For example, Das et al. (2014) reported the bacterial strain Bacillus amyloliquefaciens, isolated from chromite mine soils, exhibited a reduction rate of 2.22 mg Cr(VI) L-1 h-1 under optimized conditions using glucose as the electron sources.162 Banerjee et al. (2019) isolated a Bacillus cereus strain from an open-cast coal mine that completely reduced a 200 mg L-1 Cr(VI) concentration within 16 hours at Luria broth incubation at pH 7.5 and 37oC.205 The mechanism of microbial Cr(VI) reduction varies strain to strain and follows either or a combination of the three processes.: i) at aerobic condition, Cr(VI) reduction is associated with soluble chromate reductases that use NAD(P)H as cofactors; ii) at anaerobic condition, microbes can use Cr(VI) as an electron accepter in the electron transport chain; iii) reduction of Cr(VI) 122 also take place by chemical reactions from microbial metabolism such as amino acids, vitamins, organic acids and glutathione.6 However, pure cultures could be sensitive to environmental conditions; and keeping sterilized conditions for pure cultures is complicated for practical bioremediation.313 This leads to the question: might the problem of pure cultures be overcome by using mixed cultures which is considered more stable159? Lyu et al. (2021) reported a bacterial community YEM001 collected from sediment sludge effectively reduced Cr(VI) in landfill leachate at 2.08 mg Cr(VI) L-1 h-1 while maintaining a stable bacterial community composition during the reduction.314 Molokwane et al. (2008) reported a mixed culture of bacteria collected from a wastewater treatment plant reduced 200 mg Cr(VI) L-1 after 65 hours in aerobic cultures and 150 mg Cr(VI) mg L-1 after 130 – 150 hours under anaerobic conditions which was not achieved using purified individual species.194 Qu et al. (2018) reconstituted a mixed culture of Geotrichum sp. and Bacillus sp. isolated from a Cr(VI) contaminated site reduced 91.7% of 40 mg Cr(VI) L-1 after 7 days with 5 g L-1 glucose as electron and carbon sources at solid/liquid ratio of 1/15 (w/v) under aerobic condition, while single species could not achieve the same removal efficiency.195 So far, Cr(VI) reduction using microbial consortia provided directly by Cr-contaminated soils is seldom reported. In this study, soil samples were collected from previously reported Cr-contaminated sites in Shuitou town, Wenzhou city, China.7 Over many years of continued leather tanning activities, soil in these sites has been heavily impacted by Cr-containing tannery wastes.7 It was hypothesized that Cr-contaminated soils host microorganisms that convert the toxic Cr(VI) to nontoxic Cr(III). The objectives of this study were (i) to identify the microbial community composition in Cr-contaminated soils, and (ii) to investigate the environmental condition for Cr(VI) reduction in the presence of microorganisms provided directly by Cr-contaminated soils. 123 5.2 Materials and Methods 5.2.1 Soil sampling Soil sampling methodology and soil properties analysis methods have been described in the previous section 3.2. Four soil samples (S1-2, S1-6, S2-2 and S3-2) with different Cr concentrations were used in this study. Selected physiochemical properties of four samples are shown in Table 5.1. Native total Cr concentration was 1892 mg kg-1 in S2-2 and 3141 mg kg-1 in S3-2 (designated as high Cr-contaminated soils), indicating the accumulation of high levels of Cr in soils due to the long-term tannery waste contamination. Total Cr was 473.3 in S1-2 mg kg-1 and 296.1 mg kg-1 in S1-6, designated as low Cr-contaminated soils. Soluble Cr concentration ranged from 0.08 to 0.89 mg kg-1 in the four soil samples, demonstrating most of Cr was immobile. There was no significant difference in the pH value among the four samples that were collected from the same local large area. Table 5.1 Selected properties (means with standard deviation, n = 3) of soil S1-2, 1-6, S2-2 and S3-2. S1-2 S1-6 S2-2 S3-2 pH (H2O) 7.69 (0.06) 7.84 (0.33) 7.82 (0.39) 7.85 (0.16) Total C (%) 2.20 (0.57) 4.37 (0.12) 2.60 (0.45) 4.33 (0.37) Total N (%) 0.13 (0.01) 0.30 (0.01) 0.16 (0.01) 0.31 (0.00) Total elemental analysis K (%) 0.171 (0.008) 0.214 (0.005) 0.177 (0.005) 0.195 (0.007) Na (%) 0.021 (0.001) 0.044 (0.003) 0.033(0.009) 0.022 (0.001) Mg (%) 0.289 (0.038) 0.405 (0.026) 0.286 (0.013) 0.258 (0.003) Fe (%) 2.76 (0.11) 2.53 (0.17) 2.32 (0.07) 1.99 (0.03) Al (%) 2.41 (0.07) 2.12 (0.05) 1.98 (0.03) 1.92 (0.02) P (%) 0.059 (0.004) 0.101 (0.012) 0.061 (0.002) 0.101 (0.004) S (%) 0.026 (0.001) 0.063 (0.006) 0.066 (0.027) 0.092 (0.008) 124 Ca (%) 1.02 (0.13) 2.10 (0.37) 2.08 (0.12) 3.55 (0.05) Cr (mg kg-1) 473.3 (42.7) 296.1 (18.7) 1892 (24) 3141 (30) Mn (mg kg-1) 772.8 (5.7) 845.2 (17.5) 760.7 (63.8) 729.9 (19.5) Pb (mg kg-1) < 0.003 < 0.003 < 0.003 < 0.003 Ba (mg kg-1) 134.4 (5.6) 185.4 (15.7) 149.6 (11.9) 168.1 (3.6) Soluble elemental analysis Ca (mg kg -1) 202 (6) 265 (18) 197 (4) 176 (5) Cr (mg kg-1) 0.19 (0.00) 0.08 (0.01) 0.51 (0.01) 0.89 (0.04) 5.2.3 Soil respiration rate For estimating microbial activity, the measurement of soil respiration rate (SR) and substrate induced respiration rate (SIR) are widely used.288 SR and SIR were measured using a modified sodium hydroxide trapping method.289 The details of the measurement were shown in the previous section 4.2.7. 5.2.4 Shotgun metagenomic sequencing Traditional microbiological techniques only isolate and culture a small proportion (~1% depending on the soil types) of the microorganisms in soil samples, which poorly represents the extant microbial diversity in soils.315 In this study, shotgun metagenomic sequencing, on the basis of collective genomes that are present in soil microbiomes, was used to reveal the microbial community composition in a subset of the four samples.315 Based on soil Cr concentration, we selected a high Cr-contaminated soil S3-2 for analysis that was expected to have enriched Cr resistant and reducing microorganisms with a low Cr-contaminated soil S1-6 as a comparison. Samples were processed by Metagenome Bio Inc., Waterloo, Canada (procedure described in Appendix B). The sequencing was performed three times to get 1.0 GB depth and sequencing 125 data were then analyzed at Sangon Biotech (Shanghai, China) (procedure described in Appendix C). 5.2.2 Microbial Cr(VI) reduction experiment A batch solution experiment was conducted to determine the Cr(VI) reduction capacity of microorganisms provided by Cr-contaminated soils. Prior to use, soil samples were incubated for 5 days at a temperature at 30oC to recover the microbial activity. All the solutions were prepared using deionized water and filter-sterilized. The basic experimental unit was a 50-mL centrifuge tube containing 20 mL of K2Cr2O7 solution at a concentration of 20 mg L-1 for S2-2 and S3-2) and 10 mg Cr(VI) L-1 (for S1-2 and S1-6). Treatment 6 (T6) was added with 1.0 g non-sterilized moist Cr-contaminated soil (equivalent dry weight basis), acting as the source of soil microorganisms. Treatment 5 (T5) had 1.0 g sterilized soil into tubes while treatment 4 (T4) did not include soil as a control. The sterilized soil was generated by autoclaving the field moist Crcontaminated soil at 121°C for 30 min to destroy the soil microorganisms including those involved in Cr(VI) reduction. By comparing T6 with T4 & T5, we tested the hypothesis that microorganisms in Cr-contaminated soils would reduce Cr(VI). A volume of 0.5 mL of 100 mM Na-acetate (0.2 g L-1of Na-acetate) was added to tubes (T4 – T6) as carbon and electron sources for microbial Cr(VI) reduction. Treatments lacking added Na-acetate were run in parallel to investigate the roles of carbon and electron sources play in microbial Cr(VI) reduction (T1 – T3). For anaerobic conditions, tubes (T1 – T6) were purged with N2 gas for 90 seconds. Lids were then wrapped with Parafilm™ to decrease O2 entry. A second set of aerobic treatments were incubated without purging N2 (T7 – T12) to investigate Cr(VI) reduction by soil microorganisms would also occur aerobically. Tubes were also covered by Parafilm™ to minimize the water loss but with a hole allowing for O2 entry. A summary of the treatments was shown in Table 5.2. 126 Finally, all the tubes were incubated at 30 oC for 7 days in an incubator (VWR 1545 Digital warm air incubator) and manually shaken twice daily to ensure that soils were fully mixed with Cr(VI) solution. Three replicates were performed for each treatment. Table 5.2 A summary description of the treatments applied to K2Cr2O7 solution to determine the role of soil microorganisms in the reduction of Cr(VI) to Cr(III). Microorganisms Na-acetate (1.0 g of the four soils) (0.2 g L-1) T-1 – – + T-2 S – + T-3 Soil – + T-4 – + + T-5 S + + T-6 Soil + + T-7 – – – T-8 S – – T-9 Soil – – T-10 – + – T-11 S + – T-12 Soil + – Treatment # N2 “S” represents the addition of sterilized soils; and “Soil” resents the addition of non-sterilized soils. To determine if elevated Cr(VI) concentration inhibited microbial Cr(VI) reduction, an investigation was conducted at initial concentrations of Cr(VI) ranging from 10 – 80 mg L-1 under anaerobic and aerobic conditions. Cr(VI) reduction by soil microorganisms was also studied in the presence of various carbon and electron sources including NaHCO3, Na-citrate, sucrose, glucose, glycerol, and ethanol. 127 After 7 days, tubes were centrifuged for 15 minutes at 8200g at 20 oC (JA-25.50 fixed-angle aluminum rotor, Beckman Coulter Ireland), and each supernatant was filtered through a 0.45 µm membrane filter. Supernatants were taken and diluted for colorimetric Cr(VI) measurement according to USEPA Method 7196A.221 The remaining filtered supernatant was used for pH analysis (Thermo Scientific™ Orion 550A benchtop pH/conductivity meter). 5.3 Results 5.3.1 Active microorganisms in soils The respiration rates of soil samples are tabulated in Table 5.3. The four soils had a low respiration rate of 3.2 – 6.3 mg CO2-C kg-1d-1. Following the addition of glucose, respiration rates increased significantly. Soil S3-2 had the higher concentration of Cr (3141 mg kg-1) compared to Soil S1-6 (296 mg Cr kg-1). Interestingly S3-2 also had the higher rate both of base respiration and proportion of added C respired: 4.4% of added C respired for Soil S3-2 compared to 2.5 % for Soil S1-6. It appears that if Cr reduces soil respiration rate, the reduction occurs early and does not progress further with additional increments of Cr in soil. Although respiration rates are low, all the four soils contained active microbial populations. Table 5.3 Mean (standard deviation) of soil respiration rates during 4-day incubation (n = 3). Respiration rate (SR) in Substrate induced respiration rate Soil samples mg CO2-C kg-1 soil−1 d−1 (SIR) in mg CO2-C kg-1 soil−1 d−1 S1-2 4.8 (1.0) 10.8 (2.2) S1-6 3.2 (4.5) 15.8 (5.9) S2-2 4.7 (3.9) 26.8 (2.2) S3-2 6.3 (2.2) 28.3 (3.9) 128 5.3.2 Shotgun metagenomic sequencing results Shotgun metagenomic sequencing of S1-6 and S3-2 is summarized in Appendix Table C1. Approximately 1.0 GB of raw reads were obtained from the sequencing. After assembly, 13,638 (S1-6) and 28,888 (S3-2) contigs were gained. Sample S3-2 had more open reading frames (ORFs) when compared to S1-6, which is consistent with the distribution of contigs. The distribution of microbial taxa in soil samples at phylum, class, genus and species levels are shown in Figure 5.1 & 5.2. At the phylum level, 78 phyla are in the two soils (Figure 5.1a). The 19 phyla with relative abundance > 0.1% were summarized, accounting for 88.5% of all phyla in S1-6 and 89.0% in S3-2 (Figure 5.2a). Proteobacteria was the most abundant phylum in both samples, comprising 42.2% in S1-6 and 69.9% in S3-2 (Figure 5.2a). The Proteobacteria in S1-6 can be subdivided to Alphaproteobacteria (27.1%), Betaproteobacteria (7.6%), Gammaproteobacteria (5.5%) and Deltaproteobacteria (1.8%) (Figure 5.2b). S3-2 is rich in Alphaproteobacteria (28.1%) and Betaproteobacteria (39.1%), and poor in Gammaproteobacteria (1.77%) and Deltaproteobacteria (0.87%) (Figure 5.2b). The two soils have 1338 genera detected (Figure 5.1c). Among the genera that accounted for over 1.0% of the population in either of the two samples, Sphenodons (e.g., Sphingomonas jaspsi and Sphingomonas sp. URHD0057) were the most abundant genera, comprising 16.6% of S1-6 and 19.5% of S3-2 (Figure 5.2c).The next most abundant genus in S3-2 was Massilia (12.6%), which was a minor component (0.2%) in S1-6. The primary genera of S3-2 (Sphingomonas, Massilia, Ramlibacter, Janthinobacterium, Pedobacteria) made up 47.4% of the whole soil community. Sphingomonas (16.6%), Chloroflexi (4.5%) and Gemmatinonas (4.1%) were the three most abundant genera in S1-6. A total of 4318 species were detected from sequencing of the two soils (Figure 5.1d). Cr-resistant genera such as Desulfovibrio, Pseudomonas and Bacillus 129 appear in these Cr-contaminated soil samples (Appendix Table C3). Bacillus cereus and Bacillus subtilis, which have been previously reported as Cr(VI) reducing bacteria,165,316 occupied 0.29% of relative abundance in S3-2, whereas they accounted for less than 0.005% in S1-6 (Appendix Table C4). Zheng et al. (2015) reported that a Bacillus subtills strain BYCr-1 reduced 0.2 mM Cr(VI) to Cr(III) aerobically after 48 h incubation and the reduction mechanism was associated with a nitro/Flavin reductase NfrA that used NAD(P)H as electron donor for Cr(VI) reduction.316 Their greater abundance of Cr-resistant and reducing bacteria in S3-2 suggests a greater potential of S3-2 to reduce Cr(VI) than S1-6. Figure 5.1 Venn diagram of microbial communities at (a) phylum, (b) class, (c) genus and (d) species levels in S1-6 in blue and S3-2 in yellow. 130 Figure 5.2 Relative abundances (%) of (b) all phyla, (d) dominant classes, (c) dominant genera and (d) dominant species in S1-6 and S3-2, determined by shotgun metagenomic sequencing. At phylum level, we assigned classification confidence threshold in a lineage without classified information in the NCBI-nr database as“unclassified”.The microorganisms with a relative abundance lower than 0.10% were classified as “others”. Microbial taxa showing greater than 0.5% relative abundance are shown in class level bar graphs, while greater than 1.0% are presented in the genus and species level bar graph. 131 5.3.3 Cr(VI) reduction by soil microorganisms The extent of Cr(VI) reduction in solution by soil microbial consortia is shown in Figure 5.3. In the absence of soils, Cr(VI) concentration in solution remained unchanged regardless of Naacetate amendment and aeration status (T-1, T-4, T-7 and T10). The addition of non-sterilized high Cr-contaminated soils (S2-2 and S3-2) decreased Cr(VI) in solution to < 0.05 mg L-1 (> 99.75%) in the presence of Na-acetate under both anerobic and aerobic condition (T6 & T12). In contrast, Cr(VI) removal by non-sterilized low Cr-contaminated soils (S1-2 and S1-6) was in the range of 61.8 – 98.4% (T6 & T12). The amounts of Cr(VI) reduced in treatments with sterilized soils samples (T5 and T11) were 6.3 – 16.2%, significantly lower than that of treatments with non-sterilized soils under either oxygen conditions (T6 and T12). In the absence of Na-acetate, Cr(VI) reduction was < 11.0% (T2, T3, T8 & T9). The decrease in Cr(VI) amounts indicates the addition of Cr-contaminated soils into K2Cr2O7 solution provided the microorganisms for Cr(VI) reduction that occurs under both aerobic and anaerobic conditions, when Na-acetate is the carbon and electron source. The slight decrease of Cr(VI) content (0.7 – 2.0 mg L-1) with the addition of sterilized soils may be attributed to abiotic reduction by soil components such as organic matter264 and/or the incomplete sterilization of soil microorganisms by autoclaving.317 132 Figure 5.3 Cr(VI) concentrations in K2Cr2O7 solution amended with no soil (in black), sterilized (in red) or non-sterilized Cr-contaminated soil (in blue) after 7-day incubation with or without Na-acetate under aerobic (T7 – T12) and anaerobic (T1 – T6) conditions at 30oC using (a) low Cr-contaminated soils and (b) high Cr-contaminated soils. In each soil, means with similar letters are not significantly different (p > 0.05). 133 5.3.4 Environmental factors on microbial Cr(VI) reduction Cr(VI) removal by soil microorganisms when initial concentrations of Cr (VI) range 10 – 80 mg L-1 anaerobically and aerobically are shown in Figure 5.4. Anaerobically, S2-2 and S3-2 microorganisms achieved complete Cr(VI) removal at 20 mg Cr(VI) L-1 in 7 days. The amounts of removed Cr(VI) was highest at 40 mg L-1 with 24.8 – 29.0 mg L-1 (62.0 – 72.5%) of Cr(VI) in solution reduced. As initial Cr(VI) concentration increased, Cr(VI) reduction decreased accordingly, reducing 21.4 – 23.6 mg L-1 (35.7 – 39.4%) at 60 mg L-1 and 17.2 – 19.9 mg L-1 (21.5 – 24.9%) at 80 mg L-1, respectively. Similar to the anaerobic batch experiments, Cr(VI) removal aerobically was highest at 40 mg L-1 with 56.7 – 56.9% of Cr(VI) reduced (22.7 – 22.8 mg L-1), decreasing to 14.0 – 18.8% (11.2 – 15.1 mg L-1) when the initial Cr(VI) concentration increased to 80 mg L-1. The microbial Cr(VI) reduction aerobically was lower when compared to the anaerobic reduction at the same concentration. Cr(VI) reduction was incomplete at 20 mg Cr(VI) L-1 by S1-2 and S1-6 microorganisms, with 7.1 – 15.7 mg L-1 (anaerobic condition) and 5.3 – 6.2 mg L-1 (aerobic condition) reduced (Figure 5.4a). All the batch experiments managed a reduction of < 5.9 mg L-1 at 40 – 80 mg Cr(VI) L-1 under both oxygen conditions, indicating the Cr(VI) reduction by S1-2 and S1-6 soil microbial consortia was significantly inhibited by the increased Cr(VI) concentrations. The results also indicate Cr(VI) removal by microbial consortia in low Cr-contaminated soils is lower than in high Cr-contaminated soils in terms of the reduced Cr(VI) amount and tolerance to Cr(VI). 134 Figure 5.4 Cr(VI) removal (%) by (a)low Cr-contaminated soils (S1-2 and S1-6) and (b)high Crcontaminated soils (S2-2 and S3-2) under varying initial Cr(VI) concentration at 10 – 80 mg L-1 anaerobically (blue bar) and aerobically (yellow bar). Cr(VI) removal (%) = (initial Cr(VI) con. – measured Cr(VI) con. after incubation)/initial Cr(VI) con. × 100%. 135 Microbial Cr(VI) reduction by two high-Cr contaminated soils in the presence of the carbon source on was studied under anaerobic condition (Table 5.4). The addition of 0.2 g L-1 carbon sources provided 19.5 – 52.1 mmol L-1 e- if all the carbon can be oxidized to CO2, except NaHCO3. Cr(VI) concentrations decreased to < 1.0 mg L-1 in the presence of organic acid salts (Na-acetate, Na-citrate) and carbohydrates (D-glucose and D-sucrose), and to 2.4 – 8.0 mg L-1 when alcohols (glycerol and ethanol) were present. The concentration of Cr(VI) remained relatively steady in the presence of NaHCO3, revealing that soil microorganisms responsible for Cr(VI) reduction are not autotrophic. Table 5.4 Concentrations of Cr(VI) (mg L-1) and pH in 20 mg Cr(VI) L-1 of K2Cr2O7 solution after 7-day incubation at 30 oC in the presence of soil microbial consortia (provided by 1.0 g of S2-2 and S3-2) amended of a variety of carbon sources at 0.2 g L-1 anaerobically. S2-2 S3-2 Final pH + NaHCO3 (0 mmol L-1 e-) 18.63 (1.45) 18.09 (1.54) 7.76 (0.19) + Na-acetate (19.5 mmol L-1 e-) 0.00 (0.00) 0.00 (0.00) 7.97 (0.13) + Na-citrate (23.3 mmol L-1 e-) 0.00 (0.00) 0.84 (0.92) 8.01 (0.15) + Sucrose (28.0 mmol L-1 e-) 0.04 (0.06) 0.41 (0.58) 7.94 (0.09) + Glucose (26.6 mmol L-1 e-) 0.73 (0.23) 0.59 (0.63) 7.93 (0.10) + Glycerol (21.7 mmol L-1 e-) 7.98 (0.73) 2.98 (2.07) 7.27 (0.27) + Ethanol (52.1 mmol L-1 e-) 5.91 (1.77) 5.66 (0.53) 7.81 (0.11) 136 The dynamics of anaerobic Cr(VI) reduction by S2-2 and S3-2 soil microorganisms are shown in Figure 5.5. The amount of Cr(VI) slightly declined in the first day before dropping quickly between day 2 – 7. The reduction rate between day 2 – 7 was 4.34 ± 0.24 mg L-1 d-1 by S3-2 microbial consortium and 3.37 ± 0.54 mg L-1 d-1 by S2-2 microbial consortium. The increased supplement of Na-acetate from 0.2 – 0.5 mg L-1 did not significantly accelerate the microbial Cr(VI) reduction rate, which were 4.43 ± 0.22 mg L-1 d-1 by S3-2 microbial consortium and 3.72 ± 0.06 mg L-1 d-1 by S2-2 microbial consortium. Figure 5.5 Anaerobic Cr(VI) reduction by (a) S2-2 and (b) S3-2 soil microbial consortia at 30 oC at 20 – 60 mg Cr(VI) L-1 with the supplement of 0.2 g L-1 Na-acetate (black line) and at 40 mg Cr(VI) L-1 with 0.5 g L-1 Na-acetate (Red line) at incubation days. 137 5.4 Discussion 5.4.1 Composition of microbial communities in Cr-contaminated soils High levels of heavy metals in soils decrease microbial population, diversity and activities.318,319 In this study, soil respiration experiment was conducted to determine the active microorganisms in two contaminated soils.288 The addition of glucose to the soils resulted in significant increases in the respiration rate from 3.2 – 6.3 to 15.8 – 28.3 mg CO2-C kg-1d-1 (Table 5.4), revealing the contribution of the active microorganisms to the respiration. However, soil respiration rate even glucose-induced respiration rate in the Cr-contaminated soils were significantly lower than that in non-contaminated soils which ranged from 36.3 – 297 mg CO2-C kg-1d-1 (Figure 4.8). These results are in agreement with previous studies reporting the decreased microbial activities in soils exposed by Cr pollution.318,320,321 Shotgun metagenomic sequencing results showed that the accumulation of high levels of Cr in S3-2 was associated with an increased abundance of Proteobacteria at the phylum level and Betaproteobacteria at the class level, as compared to S1-6, with less contamination (Figure 5.2). These findings are consistent with previous reports.322–325 Sheik et al. (2012) reported that Crcontaminated soils were dominated by Proteobacteria while Actinobacteria and Acidobacteria (which were generally abundant in pristine soils) were minor components.322 Miao et al. (2015) reported the dominance of Proteobacteria for Cr(VI) wastewater treatment, with the abundance of Betaproteobacteria increased after Cr(VI) feeding.326 The shift in phylum level dominance was independent of the site and suggests that Proteobacteria and especially the class Betaproteobacteria may be the most metal tolerant microorganisms found at metal contaminated sites.322,327 138 At genus and species level, Sphingomonas (e.g., Sphingomonas jaspsi and Sphingomonas sp. URHD0057) were rich in both soils, comprising 16.6% of S1-6 and 19.5% of S3-2 (Figure 5.2). Sphingomonas are commonly regarded as generalists in soils and are able to use both labile and recalcitrant carbon sources.328 Previous studies have reported Sphingomonas causes beneficial changes including plant growth, lignocellulosic degradation and decomposition of aromatic compounds.329,328,330,331 The rich abundance of Massilia (12.6% of all genre) was observed in S32 but was a minor component (0.2% of all genera) in S1-6 (Figure 5.2). Many studies previously identified the genus Massilia in environments elevated in heavy metals such as in mine soils332,333 and serpentine soils334, suggesting that Massilia are important resistant microorganisms in heavy metal stressed environments. Members of the Bacilli genus have been reported to reduce Cr(VI) to Cr(III).165,316 Murugavelh & Mohanty (2013) reported Bacillus cereus isolated from tannery contaminated soils completely reduced 10 – 50 mg L-1 Cr(VI) in solution using glucose as carbon and electron sources.165 Zheng et al. (2015) reported a Bacillus subtilis strain isolated from rare-earth ore that reduced 0.2 mM Cr(VI) to Cr(III) in 48 h incubation.316 Liu et al. (2019) isolated 25 Cr(VI) reducing strains from Cr contaminated soils, and most were Bacilli class. In this study, Bacillus cereus and Bacillus subtilis occupied 0.29% of relative abundance in S3-2, whereas they accounted for less than 0.005% in S1-6 (Appendix Table C4). Therefore, we can conclude that the long-term Cr contamination shifts microbial composition in soils with the increasing abundance of Cr resistant and/or reducing genera such as Massilia and Bacilli. Previous studies used 15.8 GB or 18.2 GB of metagenomic sequencing data for microbial community analysis.335,336 In this study, the regular sequencing was performed three times to get the required depth to 1.0 GB in these Cr-contaminated soils with low microbial activities, which 139 would be able to give a moderate coverage of soil metagenome.337 The microbial community composition was further studied by 16S rRNA sequences amplicon sequencing338 (procedure was described in the Appendix B). Results showed Proteobacteria was the most abundant phylum in both samples, accounting for 38.3% in S1-6 and 68.4% in S3-2 (Figure 5.6). The relative abundance of Betaproteobacteria (51.3%) in S3-2 was more than 7 times higher than in S1-6 (6.8%). The Bacilli class was accounted for 2.22% of the total microbial community in S32 while only 0.30% in S-6. The 16S rRNA sequencing results were consistent with shotgun sequencing results (Figure 5.2), suggesting the 1.0 GB metagenomic data is sufficient for microbial community analysis in this study. Unexpectedly, microbial community composition in S3-2 was more diverse than in S1-6 (Figure 5.1), which was confirmed by diversity indexes (Appendix Table C2). It appears that soil diversity was reduced by fresh Cr contamination and might be recovered following long-term adaptation and vegetation activities. Giller et al. (2009) stated that increasing metal stress in soils may lead to an increase or decrease in microbial diversity, depending on the initial state of the system.319 Figure 5.6 Relative abundances (%) of (a) all phyla and (b) all classes in soil samples determined by 16S rRNA sequencing. We assigned classification confidence threshold in a lineage without classified information in the database as “unclassified”. The microorganisms with a relative abundance lower than 0.10% were classified as “others”. 140 5.4.2 Cr(VI) reduction by soil microorganisms Batch experiment results showed that the addition of 1.0 g Cr-contaminated soils decreased 10 – 20 mg Cr(VI) L-1 of 20 mL K2Cr2O7 solution in 7 days (Figure 5.3), indicating microbial consortia provided directly by Cr-contaminated soils could reduce Cr(VI) at 30oC at pH 7.8 – 8.0, with Na-acetate required as carbon and electron sources. The amounts of removed Cr(VI) were highest (29.0 mg L-1) at 40 mg Cr(VI) L-1 by soil S3-2 microbial consortia (Figure 5.4), and the performance is comparable with previously reported microbial Cr(VI) reduction studies.163,165,166,204 Various environmental factors influencing the microbial Cr(VI) reduction were evaluated in this study including initial Cr(VI) concentration, oxygen condition, and carbon source. It has been well documented that the Cr(VI) reducing capacity of soil microorganisms declined with increasing initial Cr(VI) concentration due to the toxicity at high Cr(VI).194,339 In this study, removed Cr(VI) by S1-2 and S1-6 microbial consortia was highest at 20 mg Cr(VI) L-1, with < 5.9 mg L-1 at 40 – 80 mg Cr(VI) L-1. While the amounts of removed Cr(VI) by S2-2 and S3-2 microbial consortia was highest at 40 mg Cr(VI) L-1 and decreased following increased Cr(VI) concentrations (Figure 5.4). The higher Cr(VI) removal efficiency and tolerance to Cr(VI) of S32 microorganisms compared to S1-6 coincided with the microbial composition results that showed a higher proportion of Cr-resistant and reducing bacteria in S3-2 than in S1-6 (Figure 5.2). In this study, Cr(VI) reduction by soil microbial consortia occurred under both aerobic and anaerobic conditions (Figure 5.3 – 5.4), indicating the coexistence of anerobic and aerobic or facultative microorganisms in soils.194 Anaerobic bacteria could use Cr(VI) as an electron acceptor in the electron transport chain, while aerobic Cr(VI) reduction may take place by 141 cellular reducing agents (e.g. glutathione, vitamin C) and NADH-dependent chromate reductase.140 The microbial Cr(VI) reduction under aerobic condition had a lower performance compared to the observed under anaerobic condition at the same concentration (Figure 5.4), suggesting O2 might compete with Cr(VI) as an electron acceptor.339 Han et al. (2016) reported a similar finding in which oxygen influences the metabolic rates and competes with Cr(VI) as an electron acceptor during the Cr(VI) reduction by a facultative anaerobic bacterium Shewanella MR-1.339 The lower performance of aerobic Cr(VI) reduction could also be attributed to limited microbial growth for aerobic cultures in solution at 1:10 solid/liquid ratio.195 Complete Cr(VI) removal at 20 mg L-1 by soil microorganisms when 0.2 g L-1 of Na-acetate, Na-citrate, D-glucose and D-sucrose at pH 7.8 – 8.0 are present, which is indicative of considerable metabolic diversity of soil microorganisms using a wide range of carbon sources for Cr(VI) reduction.340 Organic acids and carbohydrates have been used as carbon and energy sources in many remediation efforts that employ naturally occurring bacterial consortia.199,341 Although ethanol and glycerol have been reported to be favorable for the growth of many Cr(VI) reducing bacteria,342,343 incomplete Cr(VI) removal (60.1 – 85.1%) was observed in this study. The increased supplement of Na-acetate from 0.2 g L-1 to 0.5 g L-1 did not significantly accelerate the Cr(VI) reduction capacity (Figure 5.5), suggesting the deficiency of other resource (e.g., N, P, S) required for growth. 5.5 Summary This research showed the link between the microbial community composition of Crcontaminated soils and microbial Cr(VI) reduction. Shotgun metagenomic sequencing results demonstrated that the accumulation of high levels of Cr in a soil (e.g., 3141 mg kg-1 Cr in S3-2) shifted the microbial community composition toward increasing abundance of Cr resistant and 142 reducing microorganisms (Proteobacteria at phylum level, Betaproteobacteria at class level, and Massilia and Bacillus at genus level) when compared with a low Cr-contaminated soil. Further higher microbial Cr(VI) reduction efficiency and tolerance to Cr(VI) of consortia in high Crcontaminated soils compared to low Cr-contaminated soils was seen under in vitro conditions utilizing Na-acetate as a supplement. Combining the metagenomic and in vitro results shows that Cr contaminated soils are enriched in Cr-resistant or reducing genera, and such consortia are more resistant to Cr(VI) having a greater ability to reduce Cr(VI). Further investigations to statistically test the link are recommended. This research further demonstrated microbial reduction under both aerobic and anaerobic conditions by consortia provided directly by Cr-contaminated soils in the presence of carbon sources (e.g., Na-acetate, Na-citrate, glucose and sucrose). The ability to reduce Cr(VI) under both anaerobic and aerobic conditions suggests important metabolic diversity among consortia present in highly Cr-contaminated soils. Therefore, use of microbial consortia in high Crcontaminated soils for bioremediation of Cr(VI)-polluted wastewater could be feasible. It also confirms the remediation of Cr(VI)-polluted soils through microbial processes by adding carbon sources 341,344.341,344 However, increasing supplemental carbon from 0.2 g L-1 to 0.5 g L-1 did not significantly accelerate the microbial Cr(VI) reduction. We suggest future studies investigate the role of other resources (e.g., N, P, S, or trace nutrients) for microbial growth in order to optimize Cr(VI) reduction rate by soil microbial consortia. Further research on synergism or competition within soil microbial consortia during microbial Cr(VI) reduction is also worth investigating. 143 Chapter 6 CONCLUSIONS AND RECOMMENDATIONS 6.1 Synthesis and conclusion Environmental contamination of chromium in water and soil originates from numerous geogenic and anthropogenic activities.6,345 The fate of chromium in polluted soils is a major environmental concern to human and environmental health. Particularly, Cr(VI) has been demonstrated to be a carcinogen and have strong oxidizing capability and mutagenic properties.25 In the study area (Shuitou, China), the local environment has been heavily impacted by the disposal of Cr-containing sludge and other tannery wastes in decades of continued tanning activities since 1980s.7 It was previously reported that total Cr in collected soil samples from Shuitou was as high as 2484 mg kg-1 and leachates from these soils had up to 261 mg Cr L-1 which exceeds the regulations for drinking water.7 Mean content of Cr in collected vegetables (garlic, bokchoy, onion, radish and celery) grown on Cr-contaminated soils ranged from 4 – 48 mg kg-1. Potential bio-accumulation of Cr in humans can be derived from the transfer of Cr from soils to plants and the human consumption of vegetables. However, the speciation of Cr, especially the toxic Cr(VI), in these long-term Cr-contaminated soils was not fully understood. This dissertation revealed the complex Cr speciation and its transformation kinetics in soils and proposed the direct use of mixed microbial consortia from Cr-contaminated soils for bioremediation. Chapter 3 characterized Cr speciation (oxidation states; availability; molecular geometry) intended to highlight the genesis of immobile Cr(VI) species in long-term tannery waste contaminated soils. Chemical extraction methods showed the Cr(III) from was dominant, with toxic Cr(VI) up to 144 mg kg-1. Of the total Cr(VI) present, immobile Cr(VI) represented > 90%. 144 Synchrotron-based XANES confirmed the significant amount of immobile Cr(VI) fractions in soils. Linear combination fitting of XANES spectra revealed fractional weights (%) in samples were CrFeO3 (49.3 – 53.6), CrOOH (22.3 – 29.5) and CaCrO4 (13.2 – 25.3). The results from this study strengthen the hypothesis that Mn(III/IV) oxidation of Cr(III) was a main pathway for Cr(VI) formation in soils.245,346 Throughout 25 years of aging, the genesis of toxic Cr(VI) in Cr(III)-contaminated soils is possible under the biogeochemical conditions prevailing in these soils.253 A second finding of this work is the significant proportion of Cr(VI) species that are immobile in such naturally contaminated soils. Mechanisms such as precipitation, reabsorption and recrystallization influence the actual mobility of Cr(VI) upon weathering over periods of 25 or more years. The occurrence of immobile Cr(VI) species in long-term contaminated soils was verified by a spiking experiment over 240-day incubation. Chapter 4 documents the available Cr(VI) in soils continually decreased during aging, with immobile Cr(VI) increased by 4.5 – 31% and immobile Cr(III) increased by 68 – 95% of total spiked Cr(VI) after aging for 240 days. These results reveal a competition between reduction and immobilization of Cr(VI) in soils, in which rapid reduction prevents immobilization. Since soils with a wide range of properties were used in this study, it further reveals that Cr(VI) reduction is influenced by soil pH and total organic matter, and Cr(VI) immobilization is influenced by salt ions (such as Ca2+) and Fe oxides. From an environmental risk perspective, remediation strategies need to be applied early to ensure reduction of Cr(VI) in soils before it gets immobilized. Risks arise from the release of toxic Cr(VI) from the immobilized fraction when soil condition changes, for example, if the pH of soils decreases,347,348 or if soils are subjected to wetting-drying cycles.349 145 Numerous Cr(VI) reducing microbial strains have been isolated from various environments that have shown excellent capacity of reducing Cr(VI) to Cr(III). But it is complicated to keep sterile conditions for pure cultures during operational as distinct from laboratory bioreduction, thus the practical application could be limited. Chapter 5 documents that long-term Cr contamination shifted microbial community composition with increasing abundance of Proteobacteria phylum (69.9%), Betaproteobacteria class (39.1%), and Cr resistant and reducing genera such as Massilia (12.6%) and Bacilli (0.57%). This study also focused on the microbial Cr(VI) reduction in solution using mixed microbial consortia provided directly by Crcontaminated soils. The addition of 1.0 g Cr-contaminated soil S3-2 reduced 29.0 mg L-1 (72.5%) of 40 mg L-1 Cr(VI) in 20 mL of K2Cr2O7 solution at 30 oC, pH 7.8 – 8.0 after 7 days under anaerobic condition, when supplied with 0.2 g L-1 of Na-acetate as carbon and electron sources. Therefore, prospective application of the microbial consortia from highly Cr-contaminated soils for bioremediation of Cr(VI)-polluted environment could be expected. In summary, using samples collected from a reported contaminated site in Shuitou (China), chapter 3 revealed the complex Cr speciation in long-term contaminated soil which contained significant amount of immobile Cr(VI) species. The occurrence of immobile Cr(VI) in soils with aging was verified by a spiking experiment over 240-day incubation in Chapter 4. Environmental risks of long-term Cr-contaminated soils arise from the release of toxic Cr(VI) from the immobilized fraction. Chapter 5 of the dissertation showed Cr contaminated was associated with the increasing abundance of Cr(VI) resistant and reducing genera. In situ remediation of Cr(VI)polluted environment was achieved using microbial consortia directly provided from highly Crcontaminated soils. 146 6.2 Limitation and future research Although the immobilization of Cr(VI) in soils during aging was verified in this dissertation, the mechanism (e.g., precipitation, crystallization) still needs to be thoroughly explored by molecular-scale experiments. The influential factors affecting long-term stability of immobile Cr species needs to be investigated for appropriate management of Cr-contaminated sites and remediation project. The mechanisms how soil mixed microorganisms are involved in Cr(VI) reduction within soils requires further clarification and was not investigated in this dissertation. The competitiveness or synergy among mixed bacterial consortium also impact the efficiency of bioremediation.194,195 If the mechanisms and interspecies interactions of soil microorganisms are resolved in further studies, we will find the optimum condition for bioremediation in order to accelerate the growth of microorganisms and improve the Cr(VI) reduction efficiency. The end-product Cr(III), which ultimately determine the efficiency of bioremediation,202,350 were not characterized in this dissertation. For example, Cr(VI) reduction by Bacillus cereus XMCr-6 formed soluble Cr(III) complexes through interactions with small organic molecules and could also bind to cells by coordination with functional groups on the bacterial surfaces.203 Serratia sp. 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MicrobiologyOpen 2019, 8 (6). https://doi.org/10.1002/mbo3.762. 179 APPENDIX A Density functional theory calculations Density functional theory calculations (DFT) were performed to investigate redox reaction mechanisms, as implemented in the CP2K program package.266 DFT calculations are based on the Gaussian plane wave (GPW) scheme,351 within a generalized gradient approximation (GGA) using the Perdew–Burke–Ernzerhof exchange correlation functional.352 Plane wave density cutoff was equal to 450 Ry. Dispersion interactions were included by means of an empirical analytical potential, using the Grimme DFT-D3 method,353 with a range of 10 Å. A δ-MnO2 (101) surface sab was used to model the substrate with unit cell parameters of a = 17.554 Å, b = 30.9423 Å, c = 16.8226 Å. The deformation charge density (∆𝜌)47 describes the charge transfer, and it was calculated as: ∆𝜌 = 𝜌(Cr(III)@MnO2 ) − 𝜌(Cr(III)) − 𝜌(MnO2 ), where 𝜌(Cr(III)@MnO2 ) is the charge density of Cr(III) molecule absorbed on MnO2 surface; 𝜌(Cr(III)) and 𝜌(MnO2 ) are the charge densities of Cr(III) molecules and MnO2 surface, respectively. 180 APPENDIX B Sequencing processing procedure used by Metagenome Bio Inc. (Waterloo, Canada). Shotgun metagenomic sequencing Soil samples of 0.3 g were transferred from the received 50 ml sample tubes (kept at -20oC) to the bead tubes of Soil DNA Isolation Kit. Metagenomic DNA extraction was performed according to the supplier’s instruction. DNA concentration was fluorometrically quantified using Qubit dsDNA HS Assay Kit (ThermoFisher). Shotgun sequencing libraries were constructed with Illumina DNA Prep kit (Illumina) based on Illumina’s recommendation. Briefly, 50 ng double stranded DNA (dsDNA) was tagmented and then cleaned up. The tagmented DNA was PCR amplified for 6 cycles in order to add index adapters and sequences for sequencing. Following double-sided purification, library DNA was quantified with the Qubit dsDNA HS Assay Kit and resolved with 1.5% TAE agarose gel to estimate median size of the purified DNA. Library DNA was diluted to 4 nM with 10 mM Tris-HCl (pH 8.5), denatured and into the cartridge of MiSeq Reagent Kit v2 (2 x 250 cycles, Illumina). Sequencing was performed with MiSeq platform for three times to get the requested depth to 1.0 GB and FASTQ data of all the three runs were downloaded for genome analysis. 16S rRNA and ITS amplicon sequencing The processing procedure provided by Metagenome Bio Inc. were described as follows. Soil genomic DNA was extracted with the Sox DNA Isolation Kit according to the supplier’s recommendation. The total genomic DNA from each sample was amplified using two sets of primers: the universal archaea/bacteria primers, 515f (5’- GTGYCAGCMGCCGCGGTAA -3’) and 806r (5’GGACTACNVGGGTWTCTAAT-3’), targeting the V4 region of the archaea/bacterial 16S rRNA gene,354 and the universal fungal primers, BITS (5’- ACCTGCGGARGGATCA -3′) and B58S3 (5′-GAGATCCRTTGYTRAAAGTT-3′), targeting the ITS1 region.355 Polymerase chain 181 reaction (PCR) was prepared in triplicates in a final volume of 25 μ, containing 2.5 μl of 10 × standard Taq buffer, 0.5 μl of 10 mM deoxynucleotide (dNTP), 0.25 μl of bovine serum albumin (BSA, 20 mg ml-1), 5.0 μl of 1 μM forward primer, 5.0 μl of 1 μM reverse primer, 5.0 μl DNA, 0.2 μl of Taq DNA polymerase (5u μl-1) and 6.55 μl of PCR water. DNA was denatured at 95oC for 5 min, followed by 35 cycles of 95 oC for 30 seconds, 30 oC for 30 seconds and 72 oC for 50 seconds and then extended at 72 oC for 10 minutes. The three PCR amplicons were then pooled, and the amount of amplified target gene in the pooled PCR products for each sample was quantified on 2% TAE agarose gel. PCR products were quantified using Qubit dsDNA high sensitivity Assay Kit (Thermo Fisher Scientific Inc.) and library DNA was sequenced with MiSeq Reagent Kit v2 (2 × 250 cycles). FASTQ files was generated for taxonomic analysis. 182 APPENDIX C Sequencing data analysis by Sangon Biotech Co., Ltd. (Shanghai, China). Shotgun metagenomic sequencing Metagenomic data analysis included selection and assembly of sequencing reads, gene prediction and taxonomy assignment.338 Trimmomatic (version 0.36) was used to trim low-quality sequences, library primers and adapters.356 High quality reads were retained if satisfying the following criteria: minimum average quality score of 20 in a sliding window of 5 bp; minimum read length of 35 nt.336 The extracted reads were then aligned to NCBI NT database using BLASTn treated as contaminated reads and filtered out to output clean reads.357 Subsequently, assembly of clean reads from each sample into contigs was performed IDBAUD (version 1.1.2).358 The open reading frames (ORFs) of contigs in each sample were predicted using the metagenome implementation of Prodigal (version 2.60),359 with lengths over 100 bp translated into amino acid sequences. Clustering 95% sequence identity (90% coverage) of the whole predicted gene sequences catalog was performed using CD-HIT (version 4.6), and the longest genes of every cluster were selected to construct a nonredundant gene catalog.360 Clean reads were then mapped to the nonreductant gene catalog using Bowtie2 (version 2.1.0) and the number of reads on each gene being were counted by Samtools (version 0.1.18). Finally, gene catalogue used for further analysis were obtained and the abundance of each gene in each sample was calculated as the formula 𝑟 1 𝑘 𝑖 ∑𝑛 𝑖=1 𝐺𝑘 = 𝑙 𝑘 ∙ 𝑟 𝑙𝑖 where r was the number if reads mapped to the genes and l was the length of the gene The taxonomic annotation of predicted gene catalog was inferred by Blastp (BLAST version 2.2.21, http://blast.ncbi.nlm.nih.gov/Blast.cgi) (e-value < 1e-5, score > 60) against NCBI-nr 183 database using DIAMOND (version 0.8.20).361 The alpha indices (Shannon-Wiener and Simposon) of the soil microbial community were determined with QIIME (version 1.9.0).362 Heatmaps and Venn diagrams were created using custom R scripts. Table C1 Summary of metagenomic sequencing Category S1-6 S3-2 Total Bases Count (bp) 901,741,094 1,188,531,686 Raw reads 3,592,594 4,735,186 Clean reads 3,251,170 4,277,140 Assembled contigs 13,638 28,888 a Contig_N50 length (bp) 584 653 b Contig_N90 length (bp) 514 523 Largest contig length (bp) 2,224 28,101 Total contig length (bp) 8,231,038 20,079,457 Average Length (bp) 603.08 695.08 c 17,425 39,481 Predicted ORFs a Contig_N50 length, the length of the smallest contig in the set of largest contigs that have a combined length that represents at least 50% of the assembly.336 b Contig_N90 length, the length of the smallest contig in the set of largest contigs that have a combined length that represents at least 90% of the assembly.336 c Open reading frames, ORFs. Table C2 Microbial community Shannon-Wiener and Simpson indexes of soil samples Sample gene Shannon-Wiener Simpson S1-6 17,428 9.53 7.4e-05 S3-2 35,829 10.15 5.1e-05 184 Table C3 Genera of microorganisms in S1-6 and S3-2 with relative abundance over 0.10% Genera g__Achromobacter g__Acidimicrobiia_noname g__Acidimicrobium g__Acidobacteria_noname g__Acidobacteriaceae_noname g__Acidovorax g__Actinobacteria_class_noname g__Actinomadura g__Actinoplanes g__Adhaeribacter g__Aeromicrobium g__Afipia g__Agromyces g__Alphaproteobacteria_noname g__Altererythrobacter g__Amycolatopsis g__Anaerolineae_noname g__Arcticibacter g__Ardenticatena g__Armatimonadetes_noname g__Arthrobacter g__Azoarcus g__Azospirillum g__Bacillus g__Bacteria_noname g__Bacteroidetes_noname g__Beta_Candidatus_Accumulibacter g__Betaproteobacteria_noname g__Blastococcus g__Blastomonas g__Bordetella g__Bradyrhizobium g__Brevundimonas g__Burkholderia g__Burkholderiales_noname g__Caenimonas g__Cand_Candidatus_Entotheonella g__Cand_Candidatus_Rokubacteria_noname g__Caulobacter g__Cellulomonas S1-6 (%) 0.092 0.140 0.324 1.921 0.229 0.145 2.706 0.110 0.140 0.005 0.775 0.172 0.175 0.169 0.822 0.148 0.101 0.015 0.148 0.309 0.170 0.108 0.112 0.070 1.350 0.204 0.124 2.734 0.108 0.067 0.085 0.880 0.349 0.183 0.153 0.078 0.177 0.107 0.129 0.120 185 S3-2 (%) 0.141 0.003 0.006 0.231 0.034 0.740 0.075 0.012 0.012 1.576 0.070 0.041 0.029 0.223 0.308 0.022 0.034 0.447 0.031 0.030 0.122 0.085 0.073 0.572 0.940 0.251 0.060 1.620 0.079 0.111 0.146 0.283 1.267 0.392 0.147 1.665 0.033 0.016 0.220 0.014 g__Chlorobi_noname g__Chloroflexi_noname g__Chthoniobacter g__Collimonas g__Comamonadaceae_noname g__Comamonas g__Conexibacter g__Cupriavidus g__Curvibacter g__Cystobacter g__Dehalococcoidia_noname g__Desulfovibrio g__Dongia g__Duganella g__Dyadobacter g__Erythrobacter g__Fimbriimonas g__Flammeovirgaceae_noname g__Flavihumibacter g__Flavisolibacter g__Flavobacterium g__Frankia g__Gaiella g__Gammaproteobacteria_noname g__Gemmatimonadetes_noname g__Gemmatimonas g__Gemmatirosa g__Hassallia g__Herbaspirillum g__Herminiimonas g__Hydrogenophaga g__Hymenobacter g__Hyphomicrobium g__Ilumatobacter g__Janthinobacterium g__Jiangella g__Luteimonas g__Lysobacter g__Magnetospirillum g__Marmoricola g__Massilia g__Mastigocladus g__Mesorhizobium 0.081 4.538 0.111 0.026 0.104 0.006 1.221 0.182 0.041 0.128 0.129 0.111 0.147 0.010 0.029 0.286 0.284 0.012 0.077 0.041 0.068 0.424 0.439 1.229 0.170 4.064 3.186 0.237 0.133 0.000 0.063 0.085 0.176 0.786 0.042 0.138 0.185 0.937 0.102 0.340 0.222 0.118 0.425 186 0.145 0.379 0.017 0.202 1.147 0.150 0.132 0.274 0.321 0.054 0.015 0.022 0.031 0.986 0.217 0.232 0.005 0.121 0.137 0.124 0.147 0.021 0.076 0.061 0.005 0.302 0.358 0.141 1.178 0.128 0.233 0.201 0.039 0.033 5.585 0.010 0.023 0.134 0.048 0.020 12.616 0.260 0.226 g__Methylibium g__Methylobacterium g__Methyloceanibacter g__Micr_Candidatus_Microthrix g__Microbacterium g__Micromonospora g__Microvirga g__Mucilaginibacter g__Mycobacterium g__Myoviridae_noname g__Myxococcus g__Nitr_Candidatus_Nitrosocosmicus g__Nitriliruptor g__Nitrolancea g__Nitrososphaera g__Nitrosospira g__Nitrospira_noname g__Nocardia g__Nocardioides g__Noviherbaspirillum g__Novosphingobium g__Oxalobacteraceae_noname g__Paenibacillus g__Paraburkholderia g__Patulibacter g__Pedobacter g__Pedosphaera g__Phenylobacterium g__Phycicoccus g__Pirellula g__Planctomyces g__Polaromonas g__Pontibacter g__Pseudoduganella g__Pseudomonas g__Pseudopedobacter g__Pseudorhodoferax g__Pseudoxanthomonas g__Ralstonia g__Ramlibacter g__Reyranella g__Rhizobacter g__Rhizobium 0.197 0.182 0.104 0.591 0.312 0.176 0.209 0.027 0.782 0.000 0.144 0.111 0.160 0.105 0.132 0.179 0.026 0.184 1.726 0.044 0.419 0.016 0.161 0.065 0.321 0.096 0.355 0.167 0.285 0.200 0.119 0.098 0.020 0.015 0.292 0.009 0.016 0.151 0.071 0.305 0.239 0.192 0.169 187 0.241 0.117 0.010 0.020 0.031 0.062 0.206 0.542 0.061 0.132 0.070 0.017 0.017 0.015 0.032 0.085 0.005 0.028 0.197 0.695 0.458 0.468 0.121 0.122 0.023 4.154 0.098 0.293 0.100 0.017 0.016 0.597 0.531 0.328 0.339 0.121 0.183 0.021 0.134 5.533 0.028 0.146 0.149 g__Rhodococcus g__Rhodoplanes g__Rhodopseudomonas g__Rhodospirillales_noname g__Roseiflexus g__Rubrobacter g__Rufibacter g__Sandaracinus g__Schlesneria g__Segetibacter g__Siphoviridae_noname g__Soli_Candidatus_Solibacter g__Solirubrobacter g__Solitalea g__Sorangium g__Sphaerobacter g__Sphingobacterium g__Sphingobium g__Sphingomonadaceae_noname g__Sphingomonas g__Sphingopyxis g__Spirosoma g__Sporichthya g__Stenotrophomonas g__Steroidobacter g__Streptomyces g__Terrabacter g__Tetrasphaera g__Variovorax g__Viruses_noname 0.249 0.423 0.141 0.154 0.111 0.211 0.020 0.117 0.111 0.021 0.000 0.156 1.252 0.021 0.201 0.153 0.017 0.409 0.097 16.562 0.596 0.046 0.203 0.115 0.409 1.467 0.103 0.203 0.229 0.009 70.064 188 0.022 0.071 0.050 0.025 0.018 0.024 0.456 0.037 0.010 0.200 0.445 0.030 0.166 0.131 0.064 0.023 0.223 0.940 0.144 19.495 1.180 0.115 0.006 0.066 0.035 0.181 0.010 0.012 0.702 0.197 80.363 Table C4 Species of microorganisms in S1-6 and S3-2 with relative abundance over 0.10%. Species s__Acidimicrobium_sp._BACL27_MAG-120823-bin4 s__Acidobacteria_bacterium_DSM_100886 s__Acidobacteria_bacterium_OLB17 s__Acidovorax_delafieldii s__Actinobacteria_bacterium_IMCC26207 s__Actinobacteria_bacterium_IMCC26256 s__Adhaeribacter_aquaticus s__Aeromicrobium_sp._Root236 s__Aeromicrobium_sp._Root344 s__Aeromicrobium_sp._Root495 s__Altererythrobacter_atlanticus s__Altererythrobacter_sp._Root672 s__Altererythrobacter_troitsensis s__Arcticibacter_svalbardensis s__Arenimonas_oryziterrae s__Armatimonadetes_bacterium_OLB18 s__Bacillus_cereus s__Bacillus_subtilis s__Betaproteobacteria_bacterium_SCGC_AG-212-J23 s__Betaproteobacteria_bacterium_SG8_39 s__Betaproteobacteria_bacterium_SG8_41 s__Bradyrhizobium_elkanii s__Brevundimonas_noname s__Brevundimonas_sp._Root1279 s__Brevundimonas_sp._Root1423 s__Brevundimonas_sp._Root608 s__Caenimonas_sp._SL110 s__Candidatus_Entotheonella_sp._TSY1 s__Candidatus_Microthrix_parvicella s__Candidatus_Rokubacteria_bacterium_CSP1-6 s__Chloroflexi_bacterium_CSP1-4 s__Chthoniobacter_flavus s__Collimonas_arenae s__Comamonadaceae_bacterium_URHA0028 s__Conexibacter_woesei s__Curvibacter_sp._PAE-UM s__Cytophagales_bacterium_B6 s__Dongia_sp._URHE0060 s__Duganella_noname s__Duganella_sp._Leaf126 s__Duganella_sp._Leaf61 189 S1-6 (%) 0.108 1.454 0.355 0.019 0.964 0.478 0.005 0.197 0.128 0.125 0.106 0.581 0.064 0.015 0.124 0.191 0.005 0.000 1.268 0.979 0.301 0.113 0.118 0.079 0.046 0.052 0.078 0.129 0.591 0.107 4.224 0.111 0.004 0.076 1.221 0.006 0.000 0.147 0.003 0.000 0.000 S3-2 (%) 0.002 0.064 0.159 0.102 0.022 0.016 1.576 0.021 0.024 0.000 0.025 0.104 0.136 0.447 0.018 0.007 0.151 0.141 0.768 0.623 0.106 0.028 0.380 0.203 0.320 0.116 1.665 0.019 0.020 0.016 0.271 0.017 0.124 1.117 0.132 0.131 0.119 0.031 0.308 0.161 0.131 s__Duganella_sp._Root1480D1 s__Duganella_zoogloeoides s__Fimbriimonas_ginsengisoli s__Flammeovirgaceae_bacterium_311 s__Flavisolibacter_sp._LCS9 s__Frankia_sp._CN3 s__Hassallia_byssoidea s__Herbaspirillum_massiliense s__Herbaspirillum_sp._TSA66 s__Hydrogenophaga_noname s__Gaiella_sp._SCGC_AG-212-M14 s__Gammaproteobacteria_bacterium_SG8_30 s__Gammaproteobacteria_bacterium_SG8_31 s__Gemmatimonas_aurantiaca s__Gemmatimonas_phototrophica s__Gemmatimonas_sp._SG8_17 s__Gemmatirosa_kalamazoonesis s__Ilumatobacter_coccineus s__Ilumatobacter_nonamiensis s__Janthinobacterium_agaricidamnosum s__Janthinobacterium_lividum s__Janthinobacterium_noname s__Janthinobacterium_sp._CG23_2 s__Janthinobacterium_sp._CG3 s__Janthinobacterium_sp._HH01 s__Janthinobacterium_sp._KBS0711 s__Janthinobacterium_sp._RA13 s__Lysobacter_arseniciresistens s__Lysobacter_dokdonensis s__Lysobacter_noname s__Massilia_alkalitolerans s__Massilia_niastensis s__Massilia_noname s__Massilia_sp._9096 s__Massilia_sp._BSC265 s__Massilia_sp._JS1662 s__Massilia_sp._LC238 s__Massilia_sp._Leaf139 s__Massilia_sp._NR_4-1 s__Massilia_sp._Root335 s__Massilia_sp._Root351 s__Massilia_sp._Root418 s__Massilia_timonae 190 0.008 0.000 0.284 0.012 0.041 0.104 0.237 0.052 0.038 0.036 0.439 0.950 0.184 1.721 1.259 0.265 3.186 0.407 0.373 0.000 0.003 0.000 0.013 0.000 0.003 0.000 0.000 0.152 0.206 0.106 0.021 0.018 0.029 0.009 0.013 0.008 0.001 0.045 0.021 0.023 0.024 0.002 0.009 0.238 0.104 0.005 0.121 0.124 0.005 0.141 0.413 0.409 0.105 0.076 0.029 0.008 0.059 0.096 0.035 0.358 0.014 0.019 0.305 0.338 0.102 3.621 0.459 0.385 0.122 0.135 0.002 0.014 0.003 0.817 1.867 1.253 0.444 1.151 0.767 1.036 1.870 0.840 0.569 0.510 0.585 0.780 s__Mastigocladus_laminosus s__Methyloceanibacter_caenitepidi s__Mucilaginibacter_gotjawali s__Mucilaginibacter_paludis s__Mucilaginibacter_sp._PAMC_26640 s__Nitriliruptor_alkaliphilus s__Nitrolancea_hollandica s__Nocardioides_noname s__Nocardioides_sp._JS614 s__Nocardioides_sp._Root122 s__Nocardioides_sp._Soil777 s__Nocardioides_sp._Soil805 s__Nocardioides_sp._URHA0032 s__Noviherbaspirillum_sp._Root189 s__Oxalobacteraceae_bacterium_AB_14 s__Pedobacter_glucosidilyticus s__Pedobacter_heparinus s__Pedobacter_oryzae s__Pedobacter_sp._Hv1 s__Pedobacter_sp._V48 s__Pedosphaera_parvula s__Phenylobacterium_zucineum s__Pirellula_staleyi s__Polaromonas_sp._CF318 s__Polaromonas_sp._JS666 s__Pontibacter_actiniarum s__Pontibacter_korlensis s__Pontibacter_roseus s__Pseudoduganella_violaceinigra s__Pseudolabrys_sp._Root1462 s__Pseudopedobacter_saltans s__Pyrinomonas_methylaliphatogenes s__Ramlibacter_sp._Leaf400 s__Ramlibacter_tataouinensis s__Rhizobacter_sp._Root404 s__Rhodoplanes_sp._Z2-YC6860 s__Rhodopseudomonas_palustris s__Rhodospirillales_bacterium_URHD0088 s__Rubrobacter_xylanophilus s__Rufibacter_sp._DG15C s__Rufibacter_sp._DG31D s__Rufibacter_tibetensis s__Sandaracinus_amylolyticus 0.118 0.104 0.012 0.000 0.015 0.160 0.105 0.262 0.181 0.154 0.108 0.106 0.193 0.044 0.012 0.004 0.000 0.009 0.010 0.028 0.355 0.093 0.174 0.023 0.034 0.006 0.006 0.004 0.015 0.346 0.009 0.165 0.077 0.228 0.104 0.423 0.118 0.154 0.149 0.000 0.011 0.009 0.117 191 0.260 0.010 0.147 0.268 0.127 0.017 0.015 0.028 0.017 0.030 0.015 0.011 0.007 0.695 0.337 0.256 0.142 2.603 0.145 0.112 0.098 0.166 0.016 0.141 0.235 0.129 0.177 0.151 0.328 0.022 0.121 0.060 2.032 3.498 0.067 0.071 0.049 0.025 0.019 0.127 0.197 0.132 0.037 s__Schlesneria_paludicola s__Segetibacter_koreensis s__Solirubrobacter_soli s__Solirubrobacter_sp._URHD0082 s__Solirubrobacterales_bacterium_URHD0059 s__Solitalea_canadensis s__Sphaerobacter_thermophilus s__Sphingobacterium_sp._21 s__Sphingomonas_astaxanthinifaciens s__Sphingomonas_changbaiensis s__Sphingomonas_jaspsi s__Sphingomonas_noname s__Sphingomonas_sanxanigenens s__Sphingomonas_sp._ERG5 s__Sphingomonas_sp._KC8 s__Sphingomonas_sp._MM-1 s__Sphingomonas_sp._PR090111-T3T-6A s__Sphingomonas_sp._Root241 s__Sphingomonas_sp._Root710 s__Sphingomonas_sp._SRS2 s__Sphingomonas_sp._UNC305MFCol5.2 s__Sphingomonas_sp._URHD0057 s__Sphingomonas_wittichii s__Sphingopyxis_macrogoltabida s__Sphingopyxis_noname s__Sporichthya_polymorpha s__Steroidobacter_denitrificans s__Streptomyces_noname s__Streptomyces_thermoautotrophicus s__Tetrasphaera_sp._Soil756 s__Thaumarchaeota_archaeon_MY3 s__Variovorax_paradoxus s__Variovorax_sp._WDL1 s__[Polyangium]_brachysporum s__actinobacterium_acAcidi s__bacterium_JKG1 s__bacterium_SCGC_AG-212-C10 s__marine_actinobacterium_MedAcidi-G1 s__marine_actinobacterium_MedAcidi-G2A s__marine_actinobacterium_MedAcidi-G2B s__marine_actinobacterium_MedAcidi-G3 Sum 192 0.111 0.021 0.587 0.665 0.486 0.021 0.153 0.000 1.934 0.316 6.489 0.170 0.126 0.111 0.146 0.162 0.094 0.089 0.105 0.076 0.134 5.567 0.126 0.066 0.109 0.203 0.409 0.142 0.104 0.117 0.111 0.057 0.033 0.067 0.330 0.129 0.115 0.153 0.176 0.179 0.248 49.560 0.010 0.200 0.083 0.083 0.062 0.131 0.023 0.121 2.602 0.586 7.253 0.231 0.242 0.140 0.322 0.421 0.104 0.115 0.117 0.141 0.215 5.194 0.197 0.321 0.212 0.006 0.035 0.011 0.012 0.004 0.017 0.256 0.166 0.103 0.009 0.014 0.014 0.001 0.001 0.003 0.003 63.074 16S rRNA and ITS amplicon sequencing The raw Illumina MiSeq reads were trimmed to remove primers and barcode sequences and then quality filtered using Cutadapt (version 1.18),363 PEAR (version 0.9.8),364 and Prinseq (version 0.20.4).365 The unique sequences at or above 97% similarity were assigned into the same operational taxonomic units (OUT) and the most abundant sequence was elected as the OTU representative sequence using USEARCH (version 11.0.667). 366,367 The general taxonomic identification of the OTUs was performed using RDP classifier (16S rRNA training set 16) for 16S rRNA sequences (http://rdp.cme.msu.edu/misc/resources.jsp), and UNITE for fungal ITS sequences (http://unite.ut.ee/index.php) 368. Rarefaction curves were calculated using Mothur (version 1.43.0) based on OUT.369 The diversity index and species richness estimator (α‐diversity) for each sample, including OTU richness, coverage, Chao‐1 diversity, ACE diversity, Simpson index. and Shannon index, with respect to a sequence depth of 3%, were also calculated QIIME (version 1.9.0).362 193 A total of 6,371 (S1-6) and 22,107 (S3-2) clean reads were obtained from the 16S rRNA sequencing (Table C5) . The rarefaction curve of S1-6 and S3-2 reached the saturation phase, as well as the 99 – 100% coverage, indicating the numbers of reads sampled are sufficient.370 Table C5 Summary of species richness estimator and alpha diversity index. Sample SeqNum 16S rRNA sequencing S1-6 6371 S3-2 22107 ITS sequencing S1-6 1332 S3-2 3531 OTUs Coverage Chao1-1 ACE Simpson Shannon 469 510 0.9901 0.9983 512.4 518.1 499.8 523.9 0.0095 0.1116 5.330 3.522 125 125 0.9962 1.0000 125.2 155.0 127.3 155.0 0.1608 0.0834 3.153 3.197 OTUs – operational taxonomic units. Figure C1 Analysis of microbial community by rarefaction plot. 194