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Environmental fate and transformation of fluorotelomer compounds in landfill leachate Hamid, Hanna 2020

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ENVIRONMENTAL FATE AND TRANSFORMATION OF FLUOROTELOMER COMPOUNDS IN LANDFILL LEACHATE  by Hanna Hamid  B.Sc, Bangladesh University of Engineering and Technology, 2009 M.A.Sc, The University of British Columbia, 2013  A THESIS SUBMITTED IN PARTIAL FULFILLMENT OF THE REQUIREMENTS FOR THE DEGREE OF  DOCTOR OF PHILOSOPHY  in THE FACULTY OF GRADUATE AND POSTDOCTORAL STUDIES (Civil Engineering)  THE UNIVERSITY OF BRITISH COLUMBIA (Vancouver)   May 2020   © Hanna Hamid, 2020  ii  The following individuals certify that they have read, and recommend to the Faculty of Graduate and Postdoctoral Studies for acceptance, the dissertation entitled:  Environmental Fate and Transformation of Fluorotelomer Compounds in Landfill Leachate   submitted by Hanna Hamid in partial fulfillment of the requirements for the degree of Doctor of Philosophy In Civil Engineering  Examining Committee: Dr. Loretta Li, Department of Civil Engineering Supervisor Dr. John Grace, Department of Chemical and Biological Engineering Supervisory Committee Member Dr. Madjid Mohseni, Department of Chemical and Biological Engineering Supervisory Committee Member Dr. Susan A Baldwin, Department of Chemical and Biological Engineering University Examiner Dr. Pierre Kennepohl, Department of Chemistry University Examiner  iii  Abstract Fluorotelomer compounds (FTCs) and their transformation products are routinely detected in landfill leachate, their environmental fate and transformation in leachate are unknown. This study focused on measuring bio- and phototransformation of spiked FTCs in landfill leachate, using lab-scale experiments under environmentally relevant conditions. Spiked 8:2 fluorotelomer alcohol (FTOH) and 6:2 fluorotelomer sulfonate (FTS) and their known biotransformation products were quantified in sediment-leachate microcosms and their headspaces over 90 days. The results showed that 8:2 FTOH and 6:2 FTS persisted (half-life >>30 d) in landfill leachate-sediment microcosms. Slower biotransformation led to significant partitioning of semi-volatile 8:2 FTOH to the gas phase, suggesting that landfills may act as secondary sources for semi-volatile FTOHs in the environment. C6 – C8 and C4 – C6 perfluorocarboxylic acids (PFCAs) were the most abundant products for 8:2 FTOH and 6:2 FTS, respectively.  The effect of organic carbon and ammonia concentrations in 6:2 FTS biotransformation and PFCA formation were investigated with sediment microcosms, to which deionized water (DI) and various amounts of leachate were added. Greater biotransformation of 6:2 FTS was observed in leachate-added microcosms, compared to DI microcosms, likely reflecting substrate dependency of 6:2 FTS biotransformation. Substrate limiting conditions in DI microcosms resulted in a slightly higher formation of (C4 – C6) PFCAs compared with leachate added microcosms. To understand roles of microbial communities (e.g., heterotrophic, autotrophic) in 6:2 FTS biotransformation and PFCA production, experiments were carried out with specific substrates (i.e., glucose, ammonia) and ammonia-oxidizing inhibitor (allylthiourea) using inoculum prepared from sediment. Both heterotrophic and autotrophic bacteria were able to biotransform 6:2 FTS to varying extents. Greater biotransformation of 6:2 FTS and C4 – C6 PFCAs formation were observed in the presence of ammonia oxidizers, indicating that biological nitrification is likely to increase 6:2 FTS biotransformation and PFCA production.  Phototransformation of 6:2 FTS and PFCA production were investigated in leachate under simulated sunlight. The results showed that 6:2 FTS was undergoing indirect photolysis in leachate iv  (half-life of ∽15 days), suggesting indirect photolysis of 6:2 FTS is likely a relevant transformation pathway in sunlit aquatic environments.  v  Lay Summary Fluorotelomer compounds (FTCs) are widely used in waterproofing surface protection layers in consumer products and packaging. At the end of their useful lives, consumer products containing FTCs are disposed of at landfills. With time, FTCs are released into the landfill leachate and enter the aquatic environment through leachate disposal. FTCs can undergo transformations and produce persistent organic pollutants like perfluorocarboxylic acids (PFCAs) in leachate. Concerns exist regarding the adverse effects of FTCs and PFCAs on human and ecological health.  This study investigated the transformation processes of FTCs in landfill leachate by bacterial communities (biotransformation) and in the presence of sunlight (phototransformation). The research findings provided a better understanding of the fate of FTCs and transformation products formation in landfill leachate. The valuable insights provided by this study would allow more realistic risk assessment and development of effective exposure mitigation strategies FTCs and PFCAs.  vi  Preface I was responsible for conducting the literature review, identifying research gaps, developing the research proposal, designing and conducting experiments, sample analysis, analyzing experimental data, writing the manuscripts, and preparing the conference presentations and posters throughout the course of this research, under the supervision of Professors Loretta Li (Civil Engineering Department) and John R. Grace (Chemical and Biological Engineering Department) of the University of British Columbia, Vancouver. The microbial community analysis of landfill sediment and leachate presented in Chapters 4 and 5 were performed by Microbiome Insights (Vancouver, Canada). The following lists the manuscripts that have been published (or submitted) in peer-reviewed academic journals and presented in conferences. Peer-Reviewed Journals • An abridged version of Chapter 2 has been published in the journal Environmental Pollution. The full citation of the article is “Hamid, H., Li, L. Y., & Grace, J. R. (2018). Review of the fate and transformation of per-and polyfluoroalkyl substances (PFASs) in landfills. Environmental Pollution, 235, 74-84. https://doi.org/10.1016/j.envpol.2017.12.030” • A version of Chapter 3 has been published in the journal Science of The Total Environment. The full citation of the article is “Hamid, H., Li, L. Y., & Grace, J. R. (2020). Aerobic biotransformation of fluorotelomer compounds in landfill leachate-sediment system. Science of The Total Environment, 713, 136547. https://doi.org/10.1016/j.scitotenv.2020.136547” • A version of Chapter 5 has been published in the journal Environmental Pollution. The full citation of the article is “Hamid, H., Li, L. Y., & Grace, J. R. (2020). Formation of perfluorocarboxylic acids from 6: 2 fluorotelomer sulfonate (6: 2 FTS) in landfill leachate: vii  Role of microbial communities. Environmental Pollution, 259, 113835. https://doi.org/10.1016/j.envpol.2019.113835”  Conference Podium Presentation • Hamid, H., Li, Loretta. Y., Grace, John. R., (2019). Fluorotelomer Compounds: Curious Case of a Global Contaminant. WEST Conference, 12-14 June in Vancouver, Canada  Conference Poster Presentation • Hamid, H., Li, L. Y., & Grace, J. R. (Abstract accepted). Role of Microbial communities in Formation of Perfluorocarboxylic Acids (PFCAs) from 6:2 Fluorotelomer Sulfonate (6:2 FTS) in Landfill Leachate. SETAC Europe Annual Meeting in Dublin, Ireland, May 3-7, 2020.    viii  Table of Contents Abstract ......................................................................................................................................... iii Lay Summary .................................................................................................................................v Preface ........................................................................................................................................... vi Table of Contents ....................................................................................................................... viii List of Tables .............................................................................................................................. xiii List of Figures ............................................................................................................................. xiv List of Abbreviations ................................................................................................................. xix Acknowledgements .................................................................................................................... xxi Dedication ................................................................................................................................. xxiii Chapter 1: Introduction ................................................................................................................1 1.1 Background ..................................................................................................................... 1 1.2 Motivation for this research ............................................................................................ 2 1.3 Objectives ....................................................................................................................... 3 1.4 Research plan .................................................................................................................. 4 1.5 Thesis organization ......................................................................................................... 5 1.6 Novelty and research contribution .................................................................................. 6 Chapter 2: Literature Review .......................................................................................................8 2.1 Introduction ..................................................................................................................... 8 2.2 Poly- and perfluoroalkyl substances (PFASs) in landfill system .................................. 10  Occurrence and trends perfluoroalkyl acids (PFAAs) in landfill leachate ............... 10 2.2.1.1 Concentration and trends in North America ..................................................... 13 2.2.1.2 Concentration and trends in Europe .................................................................. 14 2.2.1.3 Concentration and trends in Australia ............................................................... 15 2.2.1.4 Concentration and trends in China .................................................................... 16  Polyfluoroalkyl compounds in landfill leachate ....................................................... 16 2.3 Release of poly- and perfluoroalkyl substances (PFASs) with landfill leachate .......... 18 2.4 Poly- and perfluoroalkyl substances (PFASs) in ambient landfill air........................... 18 2.5 Fate of poly- and perfluoroalkyl substances (PFASs) in landfill .................................. 19 2.6 Fate of PFASs in leachate treatment systems ............................................................... 22 2.7 Future research directions ............................................................................................. 22 2.8 Conclusions ................................................................................................................... 23 Chapter 3: Aerobic Biotransformation of Fluorotelomer Compounds in Landfill Leachate-Sediment........................................................................................................................................25 3.1 Introduction ................................................................................................................... 25 3.2 Materials and methods .................................................................................................. 28  Standards and reagents .............................................................................................. 28  Landfill sediment and leachate collection ................................................................. 28  Microcosm setup and sampling ................................................................................ 29  Microcosm extraction and PFAS analysis ................................................................ 30  Instrumental analysis of PFAS.................................................................................. 31  Experimental and analytical quality control and quality assurance .......................... 32 3.3 Results and discussion .................................................................................................. 32  Physical and chemical characteristics of the landfill leachate .................................. 32 ix   Microcosm operating conditions of live controls ..................................................... 33  FTCs and PFCAs in background landfill leachate and sediment ............................. 35  Biotransformation of spiked parent compounds in landfill leachate and sediment .. 36  Biotransformation products of spiked 8:2 FTOH and 6:2 FTS ................................ 36  Biotransformation rates and products yield of 8:2 FTOH and 6:2 FTS ................... 41 3.3.6.1 Comparison of 8:2 FTOH and 6:2 FTS product yields..................................... 42  Mass balances of 8:2 FTOH and 6:2 FTS ................................................................. 43  Biotransformation pathways of 8:2 FTOH and 6:2 FTS .......................................... 46 3.4 Conclusions and future research ................................................................................... 49 Chapter 4: Effect of Substrate Concentrations on Aerobic Biotransformation of 6:2 Fluorotelomer Sulfonate in Landfill Leachate ..........................................................................50 4.1 Introduction ................................................................................................................... 50 4.2 Materials and methods .................................................................................................. 51  Materials, standards and reagents ............................................................................. 51  Microcosm preparation ............................................................................................. 53  Sample collection, preparation and instrumental Analysis ....................................... 53  Poly- and perfluoroalkyl substance (PFAS) analysis................................................ 54  Organic and inorganic contents analysis................................................................... 55  DNA extraction, PCR amplification, sequencing and post-sequencing analysis ..... 55  Quality control and quality assurance ....................................................................... 55 4.3 Results and discussion .................................................................................................. 56  Microcosm monitoring.............................................................................................. 56  Microbial community analysis .................................................................................. 59 4.3.2.1 Initial microbial community composition of landfill leachate-sediment at Phylum level ..................................................................................................................... 59 4.3.2.2 Change of microbial community over 60 days ................................................. 60 4.3.2.3 Alpha diversity of the microbial community .................................................... 63  Biotransformation of 6:2 FTS and PFCA formation ................................................ 64 4.3.3.1 Biotransformation of 6:2 FTS in live-spiked microcosms ................................ 64 4.3.3.2 Formation of PFCAs in live-spiked and live-control microcosms ................... 65 4.4 Conclusions ................................................................................................................... 69 Chapter 5: Role of Microbial Communities in the Formation of Perfluorocarboxylic Acids from 6:2 Fluorotelomer Sulfonate in Leachate .........................................................................71 5.1 Introduction ................................................................................................................... 71 5.2 Materials and method .................................................................................................... 74  Sediment collection and inoculum preparation ......................................................... 74  Standards and reagents .............................................................................................. 74  Microcosm preparation ............................................................................................. 75  Sample collection, preparation and instrumental analysis ........................................ 76 5.2.4.1 PFAS analysis ................................................................................................... 76 5.2.4.2 Organic and inorganic content analyses ........................................................... 77  Quality control and quality assurance ....................................................................... 77  PCR amplification, sequencing and post-sequencing analysis ................................. 78  Data analysis ............................................................................................................. 78 5.3 Results and discussion .................................................................................................. 79 x   Microcosm monitoring and defluorination ............................................................... 79 5.3.1.1 Dissolved oxygen (DO) .................................................................................... 79 5.3.1.2 pH ...................................................................................................................... 80 5.3.1.3 Substrates .......................................................................................................... 81 5.3.1.4 Fluoride concentration ...................................................................................... 84  Microbial Community Analysis ................................................................................ 85 5.3.2.1 Initial microbial community composition ......................................................... 85 5.3.2.2 Microbial community composition at day 7 ..................................................... 86 5.3.2.3 Nitrifying bacteria ............................................................................................. 88 5.3.2.4 Alpha diversity and richness of the microbial community ............................... 89  Effect of microbial activity on 6:2 FTS biotransformation ...................................... 90  Effect of microbial activity on PFCA formation ...................................................... 92 5.3.4.1 Formation of C4 to C6 PFCAs.......................................................................... 92 5.3.4.2 Background biotransformation ......................................................................... 95 5.4 Conclusions and future research ................................................................................... 95 Chapter 6: Phototransformation of 6:2 Fluorotelomer Sulfonate in Landfill Leachate Under Simulated Sunlight ...........................................................................................................97 6.1 Introduction ................................................................................................................... 97 6.2 Materials and methods .................................................................................................. 99  Landfill leachate collection ....................................................................................... 99  Standards and reagents .............................................................................................. 99  Experimental setup and sample collection .............................................................. 100 6.2.3.1 Experimental conditions ................................................................................. 100 6.2.3.2 Light soaking chamber setup .......................................................................... 101  Instrumental analysis .............................................................................................. 102 6.2.4.1 PFAS analysis ................................................................................................. 102 6.2.4.2 Organic, inorganic and optical properties analysis ......................................... 103  Quality control and quality assurance ..................................................................... 103 6.3 Results and discussions ............................................................................................... 104  Physical and chemical and optical properties of landfill leachate .......................... 104  Effect of nitrate and humic acid on 6:2 FTS phototransformation ......................... 105  Phototransformation of 6:2 FTS in landfill leachate and product formation .......... 106 6.3.3.1 Phototransformation of 6:2 FTS ..................................................................... 106 6.3.3.2 Phototransformation products of 6:2 FTS....................................................... 108 6.4 Conclusions and future research ................................................................................. 110 Chapter 7: Conclusions and Recommendations for Future Work .......................................112 7.1 Conclusions ................................................................................................................. 112 7.2 Recommendations for future research ........................................................................ 113 References ...................................................................................................................................115 Appendices ..................................................................................................................................140 Appendix A Data on occurrence of PFASs in landfill leachate (Chapter 2) .......................... 141  Occurrence of perfluoroalkyl acids (PFAAs) in landfill leachate (concentrations expressed in ng/L) ..................................................................................................... 141  Occurrence of perfluoroalkyl sulfonamide derivatives, fluorotelomer acids in landfill leachate (concentration expressed as ng/L) ............................................................... 142 xi   Classification of landfill leachate according to age and typical characteristics (Renou et al., 2008) ................................................................................................................ 143 Appendix B Supplemental Information on Aerobic Biotransformation of Fluorotelomer Compounds in Landfill Leachate-Sediment System (Chapter 3) ........................................... 144  Standards of poly- and perfluoroalkyl substances (PFASs) ...................................... 144  Organic and inorganic contents analysis ................................................................... 147  List of chemicals used for mineral media preparation .............................................. 148  Instrumental method parameters for analysis of PFCAs and Fluorotelomer acids by LC-MS/MS ................................................................................................................ 149  Instrumental method parameters for analysis of FTOHs by GC-MS ........................ 150  Detection limits of target analytes in injection solvent (80:20/Water:Acetonitrile,v/v) 151 Figure B.2 Evolution of 7:3 FTUCA (m/z: 439>369) over time (days) in 8:2 FTOH spiked live samples. Due to lack of authentic standard, concentrations of 7:3 FTUCA has been approximated by the area counts of mass spectrometry peak of sediment-leachate extracts. ...................................................................................................................... 152 Figure B.3 Plots showing variation natural logarithm of the remaining parent compound concentrations for a) 8:2 FTOH and b) 6:2 FTS. the trend lines were used for estimating the pseudo-first order kinetics and half-lives of 8:2 FTOH and 6:2 FTS in landfill leachate-sediment. ......................................................................................... 152  Physical and chemical monitoring data for live-control microcosms (plotted in Figure 3.1) ............................................................................................................................. 153  Concentration (nmol/L) of biotransformation products of 8:2 FTOH in live-spiked microcosms (plotted in Figure 3.3) ........................................................................... 154  Concentrations (nmol/L) of 8:2 FTOH in the live-spiked microcosms (plotted in Figure 3.4) ................................................................................................................. 155  Concentrations (nmol/L) of 8:2 FTOH in the sterile-spiked microcosms (plotted in Figure 3.4) ................................................................................................................. 155  Concentration (nmol/L) of PFASs in live control microcosms (plotted in Figure 3.2) 156 Appendix C Supplemental Information on Effect of Leachate Addition on 6:2 FTS Biotransformation (Chapter 4) ................................................................................................ 157  Standards of Perfluorocarboxylic acids (PFCAs) and 6:2 FTS and their suppliers .. 157  Extraction and Clean-up of PFASs from Microcosm Extracts ................................. 158  Instrumental method parameters for analysis of PFCAs and 6:2 FTS by LC-MS/MS 159  Detection limits (ng/mL) of target analytes in injection solvent (95% aqueous methanol) ................................................................................................................... 160  Two-way ANOVA comparing Shannon indices of live-spiked microcosms on various sampling days under various treatment conditions ................................................... 162  Two-way ANOVA comparing Shannon indices of live-control microcosms on various sampling days under various treatment conditions ................................................... 163 Appendix D Supplemental Information on Role of Microbial Communities in Formation of Perfluorocarboxylic Acids from 6:2 Fluorotelomer Sulfonate (6:2 FTS) in Leachate (Chapter 5) ............................................................................................................................................. 164 xii   Relative abundances of Actinobacteria phyla under various experimental conditions after 7 days ................................................................................................................ 164  Known nitrifying bacteria identified through 16s RNA sequencing in 6:2 FTS biotransformation microcosms .................................................................................. 165  Alpha diversity of microbial communities ................................................................ 165 Chao1: ................................................................................................................................. 165 Simpson index: ................................................................................................................... 166 Shannon index: ................................................................................................................... 166  Comparing the Shannon diversity index of two samples using Hutcheson t-test (Gardener 2012). ....................................................................................................... 167  Hutcheson t-test analysis for Shannon diversity index between 0 and 7 d samples under each treatment.................................................................................................. 169  One-way ANOVA for comparing % 6:2 FTS remaining after 10 days under various treatment conditions .................................................................................................. 170  One-way ANOVA and Pairwise Comparison using Tucky test for total C4-C6 PFCA formation at day 10. ................................................................................................... 171  Comparison of experimental conditions, and major outcomes of previous and current 6:2 FTS biotransformation studies under aerobic conditions .................................... 173 Appendix E Supplemental Information on Photolysis of 6:2 Fluorotelomer Sulfonate (6:2 FTS) in Landfill Leachate Under Simulated Sunlight (Chapter 6) ...................................................... 174  Standards of per- and polyfluoroalkyl substances (PFASs) and their suppliers ....... 175  Chemicals used for phosphate buffer and stock solution preparation and their suppliers ..................................................................................................................... 175  Instrumental method parameters for analysis of PFCAs and 6:2 FTS by LC-MS/MS 178  The instrumental detection limit (µg/L) of target analytes in deionized water ......... 179  xiii  List of Tables Table 2.1 Landfill location and characteristics for leachate sampling sites ................................. 12 Table 2.2 Concentration ranges of various classes of poly- and perfluoroalkyl substances (PFASs) in ambient landfill air (pg/m3).......................................................................... 19 Table 3.1 Experimental conditions of biotransformation microcosms containing sediment, landfill leachate and mineral media ................................................................................ 30 Table 3.2 Comparison of leachate characteristics with literature values ...................................... 33 Table 3.3 Comparison of experimental conditions and major outcomes of previous and current FTOH aerobic biotransformation studies ....................................................................... 38 Table 3.4 Comparison of experimental conditions and major outcomes of previous and current 6:2 FTS biotransformation studies under aerobic condition ........................................... 39 Table 4.1 Initial characterization of sediment and leachate samples (n=3) .................................. 52 Table 4.2 Experimental conditions of 6:2 FTS aerobic biodegradation ....................................... 53 Table 5.1 Conditions of the microcosms containing alkalinity (800 mg/L CaCO3), ammonia (100 mg/L) .............................................................................................................................. 76 Table 6.1 Experimental conditions of 6:2 FTS phototransformation under simulated sunlighta 101 Table 6.2 Characteristics of filtered (0.45 µm cellulose filter) landfill leachate (n=3) used in phototransformation experiments ................................................................................. 104   xiv  List of Figures Figure 1.1 Chemical structure of fluorotelomer compounds (n=2-12, X= -OH,- SO3-) and side-chain fluorotelomer-based polymers. ............................................................................... 1 Figure 1.2 Flow chart showing research plan and outcome............................................................ 4 Figure 2.1 Poly- and perfluoroalkyl substances (PFASs) ............................................................... 9 Figure 2.2 Environmental pathways of poly- and perfluoroalkyl substances (PFASs) originating from solid wastes ............................................................................................................ 10 Figure 2.3 Concentration of perfluoroalkyl carboxylic acids (PFCAs) and perfluoroalkane sulfonic acids (PFSAs) in landfill leachate in different countries. (For studies reporting concentration from multiple landfills, median concentration has been plotted; values less than detection limit were assumed to be equal the detection limit). Note the logarithmic scale of the Y-axis. ...................................................................................... 11 Figure 2.4 Timeline (1949 to 2019) of the major production, commercialization and legislation events of perfluorocarboxylic acids (PFCAs, top) and perfluorosulfonic acids (PFSAs, bottom). Events and actions that may have resulted in increased concentrations in the environment and important findings are indicated by a red arrow. Green arrows represent phase-outs and regulatory initiatives that may result in decreased concentrations in the environment. Perfluorooctane sulfonyl fluoride (POSF) is the major raw material of perfluorooctane sulfonic acid (PFOS). Please note the uneven time scale. ....................................................................................................................... 15 Figure 2.5 Concertation of polyfluoroalkyl compounds in landfill leachate in the USA (Allred et al., 2014; Huset et al., 2011) and Canada (Benskin et al., 2012). For each study, median concentrations were calculated and plotted in a logarithmic Y axis. ............................. 17 Figure 3.1 Time variation of a) headspace oxygen content, b) pH, c) nitrogenous compounds, d) total organic carbon (TOC) in the dissolved phase of the live controls (without any spiking). Absolute range of duplicate samples (n=2) is shown by error bars................. 34 Figure 3.2 Background concentrations of a) perfluorocarboxylic acids (PFCAs) in live-controls (no 8:2 FTOH/6:2FTS spike); b) fluorotelomer acids live-controls; c) PFCAs in sterile-controls; d) fluorotelomer acids sterile-controls throughout the experimental period. Absolute ranges of the duplicate samples (n = 2) are expressed by error bars. Error bars smaller than the symbol height are not visible. .............................................................. 36 Figure 3.3 Formation of a) fluorotelomer compounds from 8:2 FTOH, b) perfluorocarboxylic acids (PFCAs) from 8:2 FTOH, c) fluorotelomer compounds from 6:2 FTS and d) short-chain PFCAs from 6:2 FTS in landfill leachate over the experimental period of 90 days. Absolute ranges of the duplicate samples (n = 2) are expressed by error bars. Error bars smaller than the symbol height are not visible. .............................................................. 40 xv  Figure 3.4 Concentrations of the 8:2 FTOH in various sampling media a) live-spiked; b) spiked sterile controls; c) concentrations of 6:2 FTS in spiked live and sterile microcosms. All concentrations are normalized to the initial spiking concentration (~500 nmol/L for 8:2 FTOH and ~3000 nmol/L for 6:2 FTS). Absolute ranges of the duplicate samples (n = 2) are expressed by error bars. Error bars smaller than the symbol height are not visible. 45 Figure 3.5  Aerobic biotransformation pathways of 8:2 FTOH and 6:2 FTS in sediment-leachate in comparison with previous studies. The metabolites in blue have been reported in previous studies .............................................................................................................. 48 Figure 4.1 Time variation of a) headspace oxygen content (%), b) pH, c) total organic carbon (TOC; mg/L), d) ammonia (mg N/L), e) nitrite (mg N/L) and f) nitrate (mg N/L) in the dissolved phase of microcosms added with deionized (DI) water, diluted leachate (50:50; leachate: DI water) and leachate. All microcosms were spiked with 6:2 FTS. The absolute differences of duplicate measurements are represented by error bars. ..... 57 Figure 4.2 Time variation of a) headspace oxygen content (%), b) pH, c) total organic carbon (TOC; mg/L), d) ammonia (mg N/L), e) nitrite (mg N/L) and f) nitrate (mg N/L) in the dissolved phase of live-control and sterile microcosms. The live-controls were added with deionized (DI) water, diluted leachate (50:50; leachate: DI water) and leachate, without any 6:2 FTS spike. The sterile microcosms were spiked with 6:2 FTS. The absolute difference of duplicate measurements is represented by the error bars. .......... 58 Figure 4.3 Relative abundance of microbial community structures at the phylum level in 6:2 FTS spiked microcosms at day 0. Lower abundance (<1%) taxa are summed and represented as “Others”. Duplicates were analyzed for 0 and 14 d samples. The absolute difference of relative abundances of duplicates samples ranged between 0.02 to 0.1%. ................ 60 Figure 4.4 Relative abundance of phylum Actinobacteria and Acidobacteria in a) live-spiked and b) live-control microcosms. Lower abundance (<1%) taxa are summed and represented as “Others”. Duplicates were analyzed for 0 and 14 d samples. The absolute difference of relative abundances of duplicates samples on 0 and 14 ranged between 0 and 0.2%. 61 Figure 4.5 Relative abundance of known nitrifying genera in a) live-spiked microcosms and, b) live-control microcosms (without 6:2 FTS spike). ......................................................... 62 Figure 4.6 Microbial diversity using Shannon index in a) live-spiked microcosms and, b) live-control microcosms (without 6:2 FTS spike). The absolute differences of Shannon index values of duplicates samples at days 0 and 14 ranged between 0.04 to 11%. ................ 63 Figure 4.7 Concentrations of 6:2 FTS in live and sterile microcosms at various sampling days, normalized to initial spiking concentration on day 0. The initial spiking concentration varied between 635 – 670 µg/L (1490 – 1570 nmol/L). The live-spiked microcosms were added with an equal volume of deionized (DI) water, diluted leachate (50:50; leachate: DI water) and landfill leachate. The absolute difference of duplicate measurements is represented with the error bars. ........................................................... 65 xvi  Figure 4.8 Formation of a) C4 – C6 perfluorocarboxylic acids (PFCAs) resulting from biotransformation of 6:2 FTS in microcosms, and b) C7 and C8 PFCAs in live-spiked microcosms added with equal volume of deionized (DI) water, diluted leachate (50:50; leachate: DI water) and landfill leachate; The initial spiking concentration of 6:2 FTS varied between 635 – 670 µg/L (1490 – 1570 nmol/L). The absolute difference of duplicate measurements is represented with the error bars. ........................................... 67 Figure 4.9 Formation of a) C4 – C6 PFCAs, and b) C7 and C8 PFCAs in live-control (no 6:2 FTS spike) microcosms added with deionized (DI) water, diluted leachate (50:50; leachate: DI water) and leachate. The absolute differences of duplicate measurements are represented with error bars. ...................................................................................... 69 Figure 5.1 Conceptual model of nitrification, showing ammonia and nitrite oxidation ............... 73 Figure 5.2 Time variation of dissolved oxygen (DO; mg/L) in a) live-spiked microcosms with heterotrophic only (HET) and heterotrophic and autotrophic (HET+AOB+NOB); b) live-spiked microcosms with nitrite-oxidizing bacteria (NOB) only and ammonia- and nitrite-oxidizing (AOB+NOB); c) sterile and live-control (no 6:2 FTS spike) microcosms with HET and HET+AOB+NOB growth; d) live-control microcosms with NOB and AOB+NOB growth. The absolute difference of duplicate measurements is represented with the error bars. ...................................................................................... 80 Figure 5.3 Time variation of pH in a) live-spiked microcosms with heterotrophic only (HET) and heterotrophic and autotrophic (HET+AOB+NOB); b) live-spiked microcosms with nitrite-oxidizing bacteria (NOB) only and ammonia- and nitrite-oxidizing (AOB+NOB); c) sterile and live-control (no 6:2 FTS spike) microcosms with HET and HET+AOB+NOB growth; d) live-control microcosms with NOB and AOB+NOB growth. The absolute difference of duplicate measurements is represented with the error bars. ................................................................................................................................. 81 Figure 5.4 Time variation of total organic carbon (TOC; mg/L) in the dissolved phase of a) live-spiked microcosms with heterotrophic only (HET) and heterotrophic and autotrophic (HET+AOB+NOB); b) live-spiked microcosms with nitrite-oxidizing bacteria (NOB) only and ammonia- and nitrite-oxidizing (AOB+NOB); c) sterile and live-control (no 6:2 FTS spike) microcosms with HET and HET+AOB+NOB growth; d) live-control microcosms with NOB and AOB+NOB growth. The absolute difference of duplicate measurements is represented with the error bars. ........................................................... 82 Figure 5.5 Time variation of ammonia (NH4+; mg N/L) in the dissolved phase of a) live-spiked microcosms with heterotrophic only (HET) and heterotrophic and autotrophic (HET+AOB+NOB); b) live-spiked microcosms with nitrite-oxidizing bacteria (NOB) only and ammonia- and nitrite-oxidizing (AOB+NOB); c) sterile and live-control (no 6:2 FTS spike) microcosms with HET and HET+AOB+NOB growth; d) live-control microcosms with NOB and AOB+NOB growth. The absolute difference of duplicate measurements is represented with the error bars. ........................................................... 83 xvii  Figure 5.6 Time variation of nitrite and nitrate (mg N/L) in the dissolved phase of a) live-spiked microcosms with heterotrophic only (HET) and heterotrophic and autotrophic (HET+AOB+NOB); b) live-spiked microcosms with nitrite-oxidizing bacteria (NOB) only and ammonia- and nitrite-oxidizing (AOB+NOB); c) sterile and live-control (no 6:2 FTS spike) microcosms with HET and HET+AOB+NOB growth; d) live-control microcosms with NOB and AOB+NOB growth. The absolute difference of duplicate measurements is represented with the error bars. ........................................................... 84 Figure 5.7 Relative abundance of microbial community structures at the phylum level in the live-spiked and live-control microcosms. Lower abundance (<0.5%) taxa are summed and represented as “Others”. (HET: heterotrophic, AOB: ammonia-oxidizing bacteria, NOB: nitrite-oxidizing bacteria) ............................................................................................... 86 Figure 5.8 Relative abundance of phylum Actinobacteria at the genus level in live-spiked and live-control microcosms. Phylum level relative abundance of Actinobacteria varied between 0.3 to 0.8% for all samples, except HET+AOB+NOB, which showed 10 – 18% Actinobacteria at day 7. (HET: heterotrophic, AOB: ammonia-oxidizing bacteria, NOB: nitrite-oxidizing bacteria). .............................................................................................. 87 Figure 5.9 Relative abundance (%) of known nitrifying bacteria at the genus level identified through 16s RNA sequencing in live-spiked microcosms. (HET: heterotrophic, AOB: ammonia-oxidizing bacteria, NOB: nitrite-oxidizing bacteria) ...................................... 89 Figure 5.10 Microbial diversity using a) Shannon index, b) Simpson index; microbial richness using c) Chao1 estimator, d) Abundance-Richness Coverage Estimator (ACE) in live-spiked and live-control microcosms. (HET: heterotrophic, AOB: ammonia-oxidizing bacteria, NOB: nitrite-oxidizing bacteria) ...................................................................... 90 Figure 5.11 Concentration of 6:2 FTS in live-spiked and sterile microcosms at various sampling days, normalized to initial spiking concentration (day 0). The initial spiking concentration varied between 750 – 800 µg/L (1756 – 1870 nmol/L). The absolute difference of duplicate measurements is represented with the error bars. (HET: heterotrophic, AOB: ammonia-oxidizing bacteria, NOB: nitrite-oxidizing bacteria). ... 92 Figure 5.12 Formation of a) C4 – C6 Perfluorocarboxylic acids (PFCAs) resulting from biotransformation of 6:2 FTS, and b) C7 to C8 PFCAs resulting from background biotransformation of unknown precursor compound in the live-spiked microcosms with heterotrophic (HET), ammonia-oxidizing bacteria (AOB) nitrite-oxidizing bacteria (NOB); The initial spiking concentration of 6:2 FTS varied between 750 - 800 µg/L in the live-spiked microcosms. The absolute differences of duplicate measurements are represented with error bars. ............................................................................................ 94 Figure 6.1 Schematic representation of the experimental setup showing the light source and sample arrangement. (See Figure E.3 for photo of light soaking chamber.) ................ 102 xviii  Figure 6.2 Percentage transmittance of UV/Vis spectrum in filtered landfill leachate and 1% 6:2 FTS solution.................................................................................................................. 105 Figure 6.3 Time variation of spiked 6:2 FTS in pH 7.1 phosphate buffer solutions in the presence of a) 25 – 154 mg/ L nitrate, b) 7 – 200 mg/L of humic acid (HA), and c) with 22 and 25 mg/L HA and nitrate. The initial spiking concentration of 6:2 FTS ranged between 110 – 130 µg/L. The errors of duplicate measurements (or triplicates for 10 mg/L nitrate) are represented by the error bars. The errors were within the range of those observed for quality control standards (±15%). ................................................................................. 107 Figure 6.4 Phototransformation of spiked 6:2 FTS in landfill leachate in irradiated and dark control solution over a period of 72 hours. The initial spiking concentration was ∽100 µg/L. The absolute differences of duplicate (and triplicates for leachate spiked conditions) are expressed by the error bars. ................................................................. 108 Figure 6.5 a) C4 and C6 perfluorocarboxylic acids (PFCAs), and b) 5:3 fluorotelomer carboxylic acid (FTCA) in landfill leachate in irradiated and dark control solution over 72 hour period. The absolute differences of duplicate are expressed as error bars. .................. 110  xix  List of Abbreviations ACE Abundance-richness coverage estimator AMO Ammonia monooxygenase AOB Ammonia-oxidizing bacteria ATU Allylthiourea D Days DI Deionized water DiPAP Disubstituted fluorotelomer phosphate esters  DNA Deoxyribonucleic acid  DO Dissolved oxygen DOC Dissolved organic carbon EtFOSAA Ethyl-perfluorooctane sulfonamide acetic aicd  EtFOSE Ethyl-perfluorooctane sulfonamidoethanol FASA Perfluoroalkane and N-alkyl perfluoroalkane sulfonamide acetic acid FOSA Perfluoroalkane and N-alkyl perfluoroalkane sulfonamide FOSE Perfluoroalkane and N-alkyl perfluoroalkane sulfonamidoethanols FTAL Fluorotelomer aldehyde FTC Fluorotelomer compound FTCA Fluorotelomer saturated carboxylic acid FTOH Fluorotelomer alcohol FTS Fluorotelomer sulfonate FTUCA Fluorotelomer unsaturated carboxylic acid H Hours HA Humic acid HDPE High-density polyethylene HET Heterotrophic bacteria LOD Limit of detection LOQ Limit of quantification MSW Municipal solid waste NOB Nitrite-oxidizing bacteria out Operational taxonomic unit PAP Polyfluorinated phosphate ester PAR Photosynthetically active radiation PCR Polymerase chain reaction PFAA Perfluoroalkyl acid PFAS Poly- and perfluoroalkylsubstance  PFBA Perfluorobutanoic acid PFBS Perfluorobutane sulfonic acid PFCA Perfluoroalkyl carboxylic acid PFHpA Perfluoroheptanoic acid PFHxA Perfluorohexanoic acid PFHxS Perfluorohexane sulfonic acid PFNA Perfluorononanoic acid PFOA Perfluorooctanoic acid  PFOS Pefluorooctane sulfonic acid PFPeA Perfluoropentanoic acid PFPeA Perfluoropentaonoic acid xx  PFSA Perfluorosulfonic acid POP Persistent organic pollutant PP Polypropylene ROS Reactive oxygen species SPE Solid-phase extraction SUVA Specific ultraviolet absorbance  WWTP Wastewater treatment plant    xxi  Acknowledgements I express my heartfelt gratitude to my supervisors and mentors, Dr. Loretta Li, and Dr. John Grace, for their excellent guidance, enduring patience, and immense support throughout the course of my studies. It would not have been possible without their continuous encouragement. I also acknowledge financial support from The Natural Sciences and Engineering Research Council of Canada, The Schlumberger Foundation through its Faculty for the Future fellowship program, and The University of British Columbia (UBC) through the Faculty of Graduate and Postdoctoral Studies. I would like to thank Dr. Madjid Mohseni, a member of my supervisory committee, for his guidance and support during my PhD program. His valuable feedback and expert advice have helped shape and refine my research. I am also deeply grateful to Dr. Laurel Schafer and Dr. Curtis Berlinguette, from UBC Chemistry Department, for allowing me to use their lab facilities. I am also grateful to Drs. Kevin Soulsbury and David Hasman from the British Columbia Institute of Technology (BCIT) for their help regarding HPLC-MS/MS analysis.  I am grateful to Matty Jeronimo, Laboratory Program Manager from the School of Population and Public Health at UBC, for his training on analytical instruments, valuable advice and support throughout my PhD. I extend my sincere appreciation to Timothy Ma, Dr. Otman Abida and Felix Shuen from the Environmental Engineering Laboratory at the Civil Engineering Department, for their help and support with experimental setup and chemical analyses.  I would like to thank Peter Edwards (PhD student in Schafer’s Lab), David Dvorak and other graduate students from Berlinguette Group for their help with lab access at the Chemistry department, also Eric Dening Jia, a former post-doc from CHBE, for helping with the experimental setup. I am immensely grateful to the landfill operators for allowing and helping me to collect landfill leachate and sediment.  Special gratitude to my husband Nasim for his patience, unconditional understanding, help and emotional support during this study; my son Ibrahim, for being understanding and patient when mommy needed to work; my parents, for believing in me and supporting me throughout my life; my friends, for their emotional support and encouragement during this journey. Without my xxii  family and friends, this long and difficult but rewarding nonetheless, journey would not have been possible.   . xxiii  Dedication    To Nasim and Ibrahim   1  Chapter 1: Introduction 1.1 Background Fluorotelomer compounds (FTCs) belong to a class of important industrial chemicals known as poly- and perfluoroalkyl substances (PFASs). FTCs are widely used in the synthesis of fluorinated surfactants and side-chain fluorotelomer-based polymers (FTPs). FTC comprises a fluorinated chain consisting of an even number of fluorinated carbons (usually between 2 and 12) attached to two nonfluorinated carbons and a functional group, such as -OH or -SO3- (Figure 1.1). Since the 1970s’, FTPs have been applied as water repellents on a wide range of finished products (e.g., textiles, apparel, carpet, fabrics), as oil and grease repellents on paper and packaging industries, and in other surface applications (e.g., paints, adhesives, waxes, polishes, metals, electronics, and industrial cleaning products) (Buck et al., 2011).   Figure 1.1 Chemical structure of fluorotelomer compounds (n=2-12, X= -OH,- SO3-) and side-chain fluorotelomer-based polymers. An increasing number of studies are showing endocrine disruptive activity of FTCs in fish (Liu et al., 2009; 2010a; Ishibashi et al., 2008) and human cells (Maras et al., 2006), as well as reproduction impairment in fish (Liu et al., 2010a) and rats (Mylchreest et al., 2005). Additionally, transformation of FTCs to perfluoroalkyl carboxylic acids (PFCAs) are well documented in various media, including activated sludge, soil, sediment (Zhao et al., 2013a, 2013b; Wang et al., 2005; 2009), atmosphere (Ellis et al., 2004), rats (Fasano et al., 2008), humans (Gomis et al., 2017) and the atmosphere (Ellis et al., 2004). PFCAs, especially the long-chain (>C7) compounds, have attracted much attention due to ubiquitous detection in the environment (Houde et al., 2011; Kato 2  et al., 2011; Kannan et al., 2004), persistence, bioaccumulation potential and adverse effects in biota and humans. This has led manufacturers to shift toward short-chain (e.g., 6:2 FTS) compounds alternatives of legacy FTC (e.g., 8:2 FTOH). Based on the historical production volume of based FTCs and certain exposure scenarios, FTCs could contribute significantly to human and wildlife exposure to PFCAs (D'eon and Mabury, 2011).  Most consumer products and packaging containing FTPs are disposed of at municipal landfills at the end of their useful lives. In many municipalities, sewage sludge from wastewater treatment plants (WWTPs), another source of FTCs, are landfilled (Guerra et al., 2014; Arvaniti et al., 2012). After a decade-long debate on the stability of FTPs, recent studies have shown that FTPs can undergo abiotic and biotic transformation under environmental conditions, (estimated half-life of 8  ̶  100 years) releasing fluorotelomer monomers (i.e., fluorotelomer alcohol (FTOH), fluorotelomer sulfonate (FTS)) (Washington and Jenkins, 2015; Rhoads et al., 2008). The fluorotelomer monomers, with part of their carbon chains perfluorinated, can readily undergo bio- and phototransformation (half-lives of < 7 days) in the environment to form transformation products such as secondary FTOH, saturated and unsaturated fluorotelomer carboxylates and PFCAs (Liu and Meija-Avendano, 2013). Based on physio-chemical properties, some anionic, water-soluble transformation intermediates can be released with landfill leachate (Yan et al., 2015; Benskin et al., 2012); neutral compounds with low water solubilities and relatively high vapour pressures (e.g., secondary FTOHs) are also released to the surrounding air (Ahrens et al., 2011). Leachate collected from engineered landfills can be treated on-site or sent to WWTPs for off-site treatment using biological systems, before their final disposal in the surface water bodies (Townsend et al., 2015). However, biological processes are not able to mineralize these classes of contaminants and act as secondary sources of FTCs and their transformation products in the aquatic and terrestrial environment (Allred et al., 2015; Eggen et al., 2010). As solid wastes have been, and will continue to be landfilled, it is, therefore, critical to investigate landfills as long-term point sources of FTCs in the environment (Li et al., 2017; Washington et al., 2015a).  1.2 Motivation for this research Degradability of FTCs has been studied using activated sludge (Zhao et al., 2013b; Wang et al., 2012a), river sediment and soil (Liu et al., 2010b), pure (Key et al., 1998) and mixed bacterial 3  culture (Liu et al., 2010a; Fromel and Knepper, 2010). The findings clearly show that transformation outcomes vary substantially with different incubating matrices for the same compound (Liu and Meija-Avendano, 2013). To date, the mode and extent of transformation of FTCs are unknown in landfills. Transformation of the widely used fluorotelomer monomers (e.g., FTOH, FTS) in landfill leachate have never been investigated to the best of our knowledge. Hence, the fate and transport FTCs released with landfill leachate in the environment is unknown. Therefore, the overall goal of the proposed research is to study the transformation of FTCs and PFCAs formation in landfill leachate under environmentally relevant conditions.  A short-chain FTC (6:2 FTS) has been chosen as the primary parent compound of interest, as 6:2 FTS is increasingly replacing legacy long-chain fluorinated compounds (Yang et al., 2014; Poulsen et al., 2011). In addition, to better understand the effect of parent compound chain length and functional group on overall fate, an aerobic biotransformation study was conducted with 6:2 FTS and 8:2 FTOH as described below. 1.3 Objectives The following specific objectives will be addressed to achieve the overall goal of this study: 1. A thorough literature review will be conducted to provide evidence of the research gaps (Chapter 2) 2. Measuring aerobic biotransformation of FTCs (i.e., 6:2 FTS and 8:2 FTOH) and quantification of known transformation products in landfill leachate under an environmentally relevant condition (Chapter 3) 3. Assessing the effect of substrate concentrations (e.g., organic carbon, ammonia) on 6:2 FTS biotransformation and PFCAs formation in landfill leachate (Chapter 4) 4. Investigating the role of bacterial communities (e.g., heterotrophic, autotrophic bacteria) in 6:2 FTS biotransformation and PFCAs formation in landfill leachate under aerobic condition (Chapter 5) 5. Investigating phototransformation of 6:2 FTS and PFCA formation in leachate under simulated sunlight (Chapter 6)  4  1.4 Research plan  Figure 1.2 Flow chart showing research plan and outcome 5  1.5 Thesis organization The thesis has been organized into seven chapters including this chapter, which provides a brief background and motivation for this research. The general and specific objectives of this study are also outlined in this chapter, together with a research plan.  In Chapter 2, a critical review of existing publications is presented summarizing the occurrence of various classes of PFASs, including FTCs, and their sources in landfills, identifying temporal and geographical trends of PFASs in landfills; delineating the factors affecting PFASs in landfills. In addition, research gaps and future research directions are outlined. Chapter 3 presents a study on the aerobic biotransformation of FTCs (i.e., 8:2 FTOH and 6:2 FTS) in landfill leachate-sediment system. The biotransformation products are identified, along with biotransformation rates and metabolite yields. The proposed biotransformation pathway is compared with previous studies. Additionally, biotransformation outcomes of 6:2 FTS and 8:2 FTOH are compared to understand the effect chain length on PFCA production. In Chapter 4, the effect of organic carbon and ammonia on biotransformation of a short-chain fluorotelomer replacement compound (6:2 FTS), was investigated in leachate. Biotransformation experiments were conducted over 60 days with sediment collected from a landfill leachate ditch, to which deionized water and various amounts of leachate were added. Persistent biotransformation products (e.g., C4 – C6 PFCAs) are also quantified in microcosms with various organic carbon and ammonia concentrations. The microbial community composition was analyzed using 16S rRNA throughout the 60 days under all experimental conditions. Chapter 5 presents the role of various microbial communities (e.g., heterotrophic, autotrophic) towards fluorotelomer compound biotransformation. Using an inoculum prepared from the sediment of a leachate collection ditch, 6:2 FTS biotransformation experiments were carried out over 10 days. Specific substrates (i.e., glucose, ammonia) and ammonia-oxidizing inhibitor (allylthiourea) were used to produce four experimental runs with heterotrophic only, ammonia and nitrite oxidizer, nitrite oxidizer only and heterotrophic with ammonia and nitrite oxidizers. The formation of PFCAs was studied under four experimental conditions. Microbial community composition was analyzed using 16S rRNA under all experimental conditions.  6  Chapter 6 investigated the aqueous photolysis of 6:2 FTS in pH 7.1 phosphate buffer solution in the presence of humic acid and nitrate under simulated sunlight. In addition, photolysis of 6:2 FTS was measured in landfill leachate under simulated sunlight. Known phototransformation products (i.e., PFCAs) are quantified as well.  Chapter 7 summarizes the major findings of the current study and their implications for the environmental fate and transformation of FTCs in landfill leachate, followed by future research recommendations. 1.6 Novelty and research contribution The novel transformation studies in leachate covered in this thesis provide valuable insight into the fate of FTCs in the environment. The biotransformation outcomes obtained under environmentally relevant conditions can be used to predict fate and transformation of FTCs during on-site and off-site treatment of leachate (e.g., evaporation pond, aerated lagoon, activated sludge system, constructed wetland), in the soil in case of subsurface contamination due to leachate migration. The transformation studies conducted with long-chain (8:2 FTOH) and short-chain (6:2 FTS) fluorotelomer compounds shed light on their relative contributions as PFCA-precursors, thereby, assisting regulatory decisions on various FTCs. A better understanding of the effect of substrates would allow predicting FTC biotransformation and PFCA formation under various environmental conditions (e.g., seasonal change, leachate disposal in surface water, etc.). The differences in microbial communities under different experimental conditions in leachate microcosms can be used to predict responses of microbial communities during bioremediation processes. Discerning the role of heterotrophic and autotrophic bacteria in 6:2 FTS biotransformation and PFCA formation would be useful to predict and optimize the performance of biological treatment systems in terms of FTCs and PFCAs removal. Knowing the expected PFCA loading and their composition (short- versus long-chain) after biological processes, could enable the design of more efficient tertiary treatment systems. Phototransformation of 6:2 FTS in leachate under simulated sunlight, not previously studied, can be used to elucidate the transformation of FTCs in sunlit leachate, as well as in the aquatic environment. Considering the significance of landfills as secondary emission sources of PFASs in the environment, the overall 7  findings of this study are crucial for risk assessment and developing exposure mitigation strategies for PFASs in the environment.  8  Chapter 2: Literature Review 2.1  Introduction Landfilling is the most common disposal method for end-of-life consumer products (Renou et al., 2008). Engineered landfills are designed to contain solid waste and collect landfill leachate while preventing the migration of the contaminants to the groundwater. Among the emerging contaminants, poly- and perfluoroalkyl substances (PFASs), detected in landfill leachate, are receiving attention due to their persistence, bioaccumulation potential and adverse effects on biota and humans (Houde et al., 2011). PFASs are a diverse group of aliphatic compounds, where all the H atoms, except those in functional groups, attached to all C atoms (perfluoroalkyl) or in at least one C (polyfluoroalkyl) have been replaced by F atoms (Figure 2.1). Due to their unique surface-active properties and high chemical and thermal stability (Buck et al., 2011), PFASs are widely used in numerous consumer products (e.g. textiles, paper, non-stick cookware, carpets, cleaning agents) and industrial applications (e.g., metal plating, fire-fighting foams, electronics production, photography) (Arvaniti et al., 2014; Kissa, 2001). There is substantial concern over the persistence, bioaccumulation potential and possible adverse effects on animals and humans of perfluoroalkyl acids (PFAAs), a class of PFASs. Among the most commonly detected PFAAs in the environment, pefluorooctane sulfonate (PFOS) has been listed under Annex B of the Stockholm Convention Treaty on persistent organic pollutants (POPs) since 2009, prohibiting its production and use, except for a few exemptions. In May 2019, perfluorooctanoic acid (PFOA) was placed on the list of substances to be eliminated under the International Stockholm Convention Treaty on POPs (Stockholm Convention, 2019, Annex A). While PFAAs may be directly released into the environment during production, usage and disposal, polyfluoroalkyl substances - the “precursors” can also be transformed abiotically or biotically into PFAAs (Figure 2.1). A variety of consumer products (e.g., paper, textiles, and carpets) and packaging containing PFAAs and their precursors are sent to municipal landfills at the end of their useful lives. In many municipalities, biosolids containing PFASs are also landfilled (Guerra et al., 2014; Arvaniti et al., 2012). Following disposal, PFASs are released from the waste through biotic and abiotic leaching (Allred et al., 2015), as shown in Figure 2.2 Depending on their physio-chemical properties, some anionic, water-soluble PFASs (e.g., PFAAs) can be released with the landfill leachate (Yan et al., 9  2015; Benskin et al., 2012); whereas neutral PFASs with low water solubilities and relatively high vapour pressures (e.g., fluorotelomer alcohols (FTOHs)) partition with landfill gas and are subsequently released to the atmosphere (Figure 2.2). Most often, leachate from lined landfills is collected and sent to wastewater treatment plants (WWTPs) for treatment before their final disposal in surface water bodies. However, WWTPs, already burdened with PFAS from wastewater, are not equipped to remove these classes of contaminants and act as secondary sources of PFASs in the aquatic environment (Allred et al., 2015; Eggen et al., 2010). Given that solid wastes have been, and will continue to be, landfilled, it is critical to investigate landfills as long-term point sources of PFASs in the environment. This is the focus of this review.  Figure 2.1 Poly- and perfluoroalkyl substances (PFASs) As an increasing number of studies are published regarding the environmental occurrence, fate and transformation of PFASs, it is important to systematically review the published literature to critically evaluate the state of the knowledge and identify the research gaps. Recent reviews of PFASs have addressed environmental biotransformation (Liu and Meija-Avendano, 2013), fate and removal of PFASs in drinking water treatment plants (Rahman et al., 2014), and WWTPs (Arvaniti and Stasinakis, 2015). A comprehensive review of the fate and transformation of PFASs in landfills is missing to date. Therefore, this study critically reviews existing publications: i) To identify the temporal and geographical trends of PFAS occurrence in the landfill; ii) To summarise the fate and transformation of PFASs in the landfill; iii) To identify research gaps and future research directions.  10   Figure 2.2 Environmental pathways of poly- and perfluoroalkyl substances (PFASs) originating from solid wastes 2.2 Poly- and perfluoroalkyl substances (PFASs) in landfill system  Occurrence and trends perfluoroalkyl acids (PFAAs) in landfill leachate High variabilities in PFAAs profiles (Figure 2.3) in landfill leachate have been reported in studies conducted in North America, Europe, China and Australia (Yan et al., 2015; Clarke et al., 2015; Allred et al., 2014; Bossi et al., 2008; Kallenborn et al., 2004), likely reflecting the variability in dominant PFAS products and their manufacturing methodologies (Gallen et al., 2016). To date, no occurrence data exist for landfills in developing countries. PFCAs, detected in the µg/L range, are generally found to be the dominant PFAS (Fuertes et al., 2017; Allred et al., 2014; Li et al., 2012; Huset et al., 2011), with short-chain (C4  ̶  C7) PFCAs being more abundant than long-chain (≥C8) PFCAs (Fuertes et al., 2017; Li et al., 2012; Busch et al., 2010; Bossi et al., 2008; Kallenborn et al., 2004). The dominance of short-chain PFCAs is probably due to their preferential release and leaching from municipal solid waste (MSW), consistent with their higher aqueous solubilities and lower organic carbon–water partition coefficients relative to long-chain homologues (Yan et al., 11  2015). Studies also reported ∑PFAAs in leachate from landfills, closed 2 – 4 decades ago, in the range of hundreds to a few thousand ng/L (Gallen et al., 2016; Allred et al., 2014; Huset et al., 2011). The geographical and temporal trends of PFAAs are discussed below considering the published concentration ranges as (Figure 2.3) and landfill site characteristics compiled in Table 2.1.   Figure 2.3 Concentration of perfluoroalkyl carboxylic acids (PFCAs) and perfluoroalkane sulfonic acids (PFSAs) in landfill leachate in different countries. (For studies reporting concentration from multiple landfills, median concentration has been plotted; values less than detection limit were assumed to be equal the detection limit) (Fuertes et al., 2017; Yan et al., 2015; Allred et al., 2014; Gallen et al., 2014; Benskin et al., 2012; Busch et al., 2010; Woldegiorgis et al., 2006). Note the logarithmic scale of the Y-axis.   110100100010000Sweden Germany Spain USA Canada China AustraliaConcentration (ng/L)PFBAPFPeAPFHxAPFHpAPFOAPFNAPFDAPFBSPFHxSPFOS12  Table 2.1 Landfill location and characteristics for leachate sampling sites *WWTP: wastewater treatment plant; MBR: membrane bioreactor; RO: reverse osmosis; UF; ultrafiltration; NF; nanofiltration; PFAS: per-and polyfluoroalkyl compounds; N/A: not available  Reference Woldegiorgis et al., (2008) Huset et al., (2011) Benskin et al., (2012) Busch et al., (2012) Perkola et al., (2013) Allred et al., (2014) Yan et al., (2015) Gallen et al., (2016) Fuertes et al., (2017) Location (Country) Strandmossen, Djupdalen (Sweden) Gulf Coast, Pacific Northwest, west coast, Mid Atlantic states, Southeast (USA) Pacific Northwest (Canada) (Germany) Espoo (Finland) (USA) Changzhou, Guangzhou, Nanjing, Shanghai, Suzhou (China) (Australia) (Spain) Landfill Sites 3 4 active (since 1996), 2 closed (operated during 1982-1993) 2 active since 1960's 8 closed, 14 active 1 closed (1987-2007) 5 active (since 1990's), 1 closed (1975-1990) 4 active, 3 closed 6 active, 8 closed) 2 active, 2 closed between 2015 and 2015 PFAS Analytes 13 24 24 43 4 70  14 14 16  Leachate System N/A recirculation, except one flow-through system flow-through and recirculated flow-through N/A N/A flow-through  flow-through except one recirculated  flow-through  Sampling Year November 2005 2006 February-June, 2010 not available October 2009 and June 2010 N/A Spring, 2013 February-April, 2014 March 2015 Waste Type N/A primarily municipal primarily municipal municipal and commercial primarily municipal municipal and commercial municipal primarily municipal and commercial primarily municipal Sampling Method grab samples  grab samples  grab sample  N/A 24-hr composite  grab samples  grab sample from lift station grab sample from sump  grab sample Leachate Treatment System aerobic pond N/A off-site at WWTP biological and physical  N/A N/A off-site with a two-stage process (MBR/RO or NF) off-site at WWTP, off-site with a two-stage process (MBR/UF) Estimated PFAS loading in leachate N/A N/A 8-25 kg/y/landfill N/A N/A N/A 3100 - 4000 kg/y (based on total nationwide leachate generation) N/A 1 kg/y/landfill 13  2.2.1.1 Concentration and trends in North America PFAAs in leachate from landfills in the USA have been studied by Huset et al. (2011) and Allred et al. (2014) (data in A.1 of Appendix A). Both studies found PFCA to be the dominant of all PFAS class, contributing 20 – 90% of the ∑PFASs (molar concentration basis), with concentrations ranging between 10 and 8900 ng/L (Allred et al., 2014; Huset et al., 2011). In general, the abundance of ≤C7 PFCAs was reported (Figure 2.3), consistent with the production shift towards shorter chain perfluorinated compounds and the observed increase of PFBA in consumer products in the USA in the 2000s (Liu et al., 2014). The high frequency of PFOA detection in consumer products (Vestergren et al., 2015; Liu et al., 2014), along with historical use (Figure 2.4) explain the high PFOA concentration (often comparable to shorter chain PFCAs) observed in landfill leachate (Allred et al., 2014; Huset et al., 2011). While the median concentration reported by Allred et al., (2014) exceeded 1000 ng/L for all C4 – C8 PFCAs, Huset et al., (2011) observed smaller concentrations (100 – 600 ng/L) for the same compounds. The observed difference in concentrations could be a result of variation in waste composition. For example, Lang et al., (2017) observed short-chain (≤C6) PFAS to be dominant in live anaerobic landfill reactors with carpets, as opposed to the live reactors with clothing that accumulated short-chain PFCAs, PFOA, and 8:2 fluorotelomer sulfonate. Also, two live reactors containing clothing showed high variability in total PFAS release, partly due to the uneven presence of PFASs in the clothing (Lang et al., 2016). This indicates that variation in the waste composition may account for some of the observed differences in the PFAS trend. Age of the landfill and leachate management system (i.e., leachate recirculation vs flow-through), as described in Table 2.1, could be other contributing factors. Perfluorosulfonic acids (PFSAs) concentrations in leachate have varied in the range of 50  ̶  3200 ng/L in the USA, with median concentrations being of a few hundred ng/L for PFBS, PFHxS and PFOS (Allred et al., 2014; Huset et al., 2011) as shown in Figure 2.3. While PFOS was detected in all leachate samples, its concentration was generally lower than that of PFBS and PFHxS (Allred et al., 2014; Huset et al., 2011). This dominance of shorter chain PFSAs over historically used PFOS could be indicative of the transition towards C4-based chemistry after 2002 as shown in Figure 2.4 (Vestergren et al., 2015; Huset et al., 2011; Lindstrom et al., 2011). Leachates from waste cells closed in 1993 or earlier also show dominance of PFBS and PFHxS, possibly indicating 14  existence of C4-based chemistry prior to 2002 (Huset et al., 2011), in addition to the higher mobility of shorter chain PFSAs, leading to their release in aqueous phase (Higgins and Luthy, 2006). PFCA concentrations in leachates collected from Canadian landfills have been in the range of tens to few thousands of ng/L, with ≤C8 PFCAs (especially, PFPeA and PFHxA) being more abundant (Li et al., 2012; Benskin et al., 2012), similar to the trend in the USA. PFSAs, namely PFBS, PFHxS and PFOS, also varied within the same range, and the median PFHxS concentration (200 ng/L) was shown to be higher than for PFOS and PFBS in a Cross-Canada study (Li et al., 2012). Landfill gas condensate has been reported to contain C4  ̶  C8 PFAAs, with PFBS being the dominant compound at concentration 1000 ng/L (Li, 2011). 2.2.1.2 Concentration and trends in Europe PFAAs have been reported in several EU countries including Spain (Fuertes et al., 2017), Finland (Perkola and Sainio, 2013), Norway (Knutsen et al., 2019; Kallenborn et al., 2004), Sweden (Woldegiorgis et al., 2006), Denmark (Bossi et al., 2008), and Germany (Busch et al., 2010). Most of these studies (except for Knutsen et al., 2019, Fuertes et al., 2017 and Busch et al., 2010) included fewer than 10 PFAAs with concentration reported to be in the range of <1 – 1800 ng/L. The median concentrations of all PFAAs were below 550 ng/L, showing a narrower range compared to North America. Higher abundances of PFOS and PFOA compared to shorter chain PFAAs were also observed. More recent studies, including more than 10 PFAAs, in Spain (Fuertes et al., 2017) and Germany (Busch et al., 2010), reported higher abundances of shorter chain PFAAs (≤C7), a trend similar to that in North America. This is likely caused by the phase-out of C8 based chemistry by major manufacturers and the ban on PFOS by EU in 2006 (Figure 2.4). 15   Figure 2.4 Timeline (1949 to 2019) of the major production, commercialization and legislation events of perfluorocarboxylic acids (PFCAs, top) and perfluorosulfonic acids (PFSAs, bottom). Events and actions that may have resulted in increased concentrations in the environment and important findings are indicated by a red arrow. Green arrows represent phase-outs and regulatory initiatives that may result in decreased concentrations in the environment. Perfluorooctane sulfonyl fluoride (POSF) is the major raw material of perfluorooctane sulfonic acid (PFOS). Please note the uneven time scale. Adapted from (Land et al., 2015; Lindstrom et al., 2011) 2.2.1.3 Concentration and trends in Australia There has been only one study (Gallen et al., 2016) that reported the occurrence of PFAAs in leachate from 6 active and 8 closed landfill sites in Australia, as shown in Table 2.1. The PFCA and PFSA concentration ranged from <0.5 to 5700 ng/L and <0.5 to 1900 ng/L, respectively. While these ranges are slightly smaller than for the USA, the median concentrations were <550 ng/L for all PFAAs, except PFHxA (970 ng/L), and are similar to those in Europe (Gallen et al., 2016). The PFAA trend was also similar to the observed trends in other countries, with ≤C7 PFAAs being more abundant. In general, Gallen et al., (2016) reported overall higher concentrations of PFAAs (especially PFHxA and PFHpA) in leachate from active MSW landfills compared to closed ones. Interestingly, a landfill containing construction and demolition waste closed in 2011 had higher 16  PFAAs than active MSW landfills, indicating the significance of the type of waste (Gallen et al., 2016).  2.2.1.4 Concentration and trends in China PFAAs have been reported in leachates from active and closed MSW landfills in China (Qi et al., 2018; Yan et al., 2015) with concertation range from 70 to 214,000 ng/L for PFCAs (C4-C8) and 30 to 416,000 ng/L for PFSAs (Figure 2.3). While these ranges are orders of magnitudes higher than for other countries, the median concentration for most PFAAs (e.g., PFBA, PFOS, PFOA, PFBS, PFPeA) in the range of several thousand ng/L was also higher than the values reported in North America, Europe and Australia. In contrast to other studies, PFOA (mean contribution 29%) was the most abundant PFAAs, followed by PFBS (26%) and PFPrA (16%). This finding is not unexpected, considering that China is one of the few remaining major producers and the largest consumer of PFOA and its salts in the world (Li et al., 2015a). Despite the high abundance of PFBS in leachate, suggesting a shift towards C4-based chemistry, high PFOS concentrations (1000  ̶  6000 ng/L) were reported in Chinese landfill leachates (Yan et al., 2015). It is noteworthy that, following the phase-out of PFOS by its largest manufacturer (3M) in the USA, production in China grew rapidly (Figure 2.4) from 50 tons/year in 2004 to approximately 200 tons/year by 2006, half of which was manufactured for export (Yang et al., 2014). Since the addition of PFOS to Annex B of the Stockholm Convention in 2009, its manufacture has been banned in China. Nonetheless, past manufacturing history could be a key factor behind the observed relatively high concentration of PFOS. Other factors could be due to (1) long use lifetimes of major PFOS-containing products (e.g., carpets, textiles); (2) long residence of PFOS-containing MSW in the landfill and/or (3) on-going uses of PFOS-containing products in China (Yan et al., 2015).  Polyfluoroalkyl compounds in landfill leachate Polyfluoroalkyl compounds and other precursors of PFAAs and their transformation intermediates have been studied to a lesser extent than PFAAs in landfills. To date, n:2 fluorotelomer carboxylic acids (n:2 FTCAs), unsaturated fluorotelomer carboxylic acids (n:2 FTUCAs), n:3 fluorotelomer carboxylic acids (n:3 FTCAs), fluorotelomer sulfonates (n:2 FTSs), and perfluoroalkane sulfonamidoacetic acids (FASAAs, MeFASAAs, and EtFASAAs) have been detected in landfill leachate (see A.2 of Appendix A). Recent studies suggest that intermediates of FTOHs (i.e., 17  FTCAs and FTUCAs) are often present at concentrations (Figure 2.5) comparable to PFAAs (hundreds of ng/L range). Benskin et al., (2012) studied temporal trends of 14 PFAAs and 10 PFAA-precursors in a municipal landfill in Canada over a 5-month period. The authors observed strong correlations between several PFAAs and their precursors (e.g., PFOA with 6:2 and 8:2 FTCAs; PFNA, perfluorodecanoic acid (PFDA) with 8:2 and 10:2 FTCA). Two other studies (Yan et al., 2015; Busch et al., 2010) also reported increases in PFAA concentration after aerobic biological treatment, possibly due to transformation of PFAA-precursors (discussed in section 2.6) (Yan et al., 2015; Busch et al., 2010). Similarly, Allred et al., (2015) reported n:3 FTCAs, transformation intermediate of FTOH under anaerobic conditions, to be the second-largest contributor to the total loading of PFASs in a landfill leachate monitoring study.   Figure 2.5 Concertation of polyfluoroalkyl compounds in landfill leachate in the USA (Allred et al., 2014; Huset et al., 2011) and Canada (Benskin et al., 2012). For each study, median concentrations were calculated and plotted in a logarithmic Y axis. 110100100010000FOSAFOSAAMeFBSAAMeFPeSAAMeFHxSAAMeFHpSAAMeFOSAAEtFBSAAEtFPeSAAEtFHxSAAEtFHpSAAEtFOSAA6:2FTCA8:2FTCA10:2FTCA3:3 FTCA5:3 FTCA7:3 FTCA6:2 FTUCA8:2 FTUCA10:2 FTUCA4:2 FTS6:2 FTS8:2 FTSFASAAs and N- alkyl FASAAs n:2 FTCAs and n:2 FTUCAsn:2 FTSsConcentration (ng/L) Huset et al., (2011) USABenskin et al., (2012)CanadaAllred et al.,  (2014) USA18  2.3 Release of poly- and perfluoroalkyl substances (PFASs) with landfill leachate The estimated annual mass budget of PFASs based on measured PFAS concentration and yearly leachate flow ranges from <1 to 25 kg/yr per landfill as shown in Table 2.1 (Fuertes et al., 2017; Yan et al., 2015; Benskin et al., 2012; Busch et al., 2010; Oliaei et al., 2006). The estimated release is in a similar range to the ∑PFASs discharge from WWTPs, which was reported to be tens to hundreds of g/d (3 – 30 kg/y) (Ahrens, 2011). However, release data should be interpreted with caution, as these values are highly dependent on the leachate quantity (which depends on the size and age of the landfill, precipitation on-site, etc.) and the number of analytes included in the study. As discussed in section 2.2.2, PFAA precursor concentrations could be comparable to or even higher than for the PFAAs. Therefore, measurement of only PFAAs could grossly underestimate the total release of PFASs from landfills. 2.4 Poly- and perfluoroalkyl substances (PFASs) in ambient landfill air The role of landfills as PFASs sources of gaseous emission to the atmosphere has received little attention. Two studies (Ahrens et al., 2011; Weinberg et al., 2011) have reported higher (2  ̶  30 times) PFASs in landfill ambient air compared to control sites that are presumably not contaminated with landfill emissions. Ambient landfill air predominantly contained FTOHs with a concentration range in ng/m3, being >90% of total precursor compounds measured (Table 2.2). While FOSAs, FOSEs were also detected, their concentrations in the pg/m3 range were orders of magnitude lower than for the FTOHs (Ahrens et al., 2011; Weinberg et al., 2011). 8:2 FTOH was found to be the highest sole contributor (50 - 65% of the ∑FTOHs, FOSAs, FOSEs), followed by 6:2 FTOH (15 - 40%) FTOHs (Ahrens et al., 2011; Weinberg et al., 2011). The higher abundance of 8:2 FTOH over 6:2 FTOH has been reported (Jahnke et al., 2007; Shoeib et al., 2006) to be typical for urban air, which is also supported by a recent survey (Vestergren et al., 2015) of PFASs in consumer products (imported from China) in Norway, which showed that 6:2 and 8:2 FTOHs were the most abundant extractable PFASs.     19  Table 2.2 Concentration ranges of various classes of poly- and perfluoroalkyl substances (PFASs) in ambient landfill air (pg/m3) PFAAs were also detected in the particulate phase (Weinberg et al., 2011) and gas phase (Ahrens et al., 2011) of ambient landfill air. PFBA, PFHxA, and PFOA were detected most frequently and at higher concentrations compared to other PFAAs in the gas phase (Ahrens et al., 2011). This might indicate an abundance of PFOA and shorter-chain PFCAs in landfill waste or reduced availability of longer-chain PFCAs in the air due to their higher affinity towards solid particles (Arvaniti et al., 2012). This dominance of the short and even chain length PFCAs is also consistent with the PFCA distribution in landfill leachates from 22 sites in Germany (Busch et al., 2010) indicating that this pattern is typical for landfill emissions. Although PFOS is frequently detected in landfill leachate, it exhibited very low air concentrations at the landfill sites (<5 pg/m3) likely due to strong sorption of PFOS to landfill solids; efficient trapping of PFOS in landfill gas collection; and partitioning of PFOS to landfill leachate (Ahrens et al., 2011). 2.5 Fate of poly- and perfluoroalkyl substances (PFASs) in landfill Following landfilling, PFASs undergo long-term leaching, as well as the transformation of the precursor compounds, processes that are affected by the physio-chemical properties of the PFASs, as well as the landfill leachate (Yan et al., 2015). As the landfilled waste passes through successive stages of aerobic, acetogenic, methanogenic stabilization stages, significant changes occur in the physio-chemical properties, such as pH and organic and inorganic constituents (shown in A.3 of Appendix A) of the leachate (Renou et al., 2008), likely affecting the mobility and transformation of PFASs. In most cases, leachate from various waste cells undergoing varying states of decomposition is collected together and subsequently sampled for PFAS analysis, which ∑PFAAsa ∑FTOHsb ∑FTCAc ∑FOSAs, FOSEsd      ∑PFAS WWTP /∑PFASreference site Landfill status Reference 130-320 2500-26000 Not measured 60-120 5 to 30  active (Ahrens et al., 2011)  5-10e 70-100 1-10 6-20 1.5 to 2.5 Closed for last 6 years (Weinberg et al., 2011) <DLf-40e 120-660 <DL-20 7-20 1.5 to 3 active (Weinberg et al., 2011) aperfluorocarboxylic and perfluorosulfonic acids; bfluorotelomer alcohols; cfluorotelomer acids; dperfluoroalkane and N-alkyl perfluoroalkane sulfonamide and sulfonamidoethanols; emeasured in particulate phase; fdetection limit;  20  represents a significant challenge in terms of understanding the effect of waste evolution on PFAS release and transformation inside a landfill.  Benskin et al., (2012) monitored patterns of 24 PFASs (14 PFAA and 10 PFAA-precursors), in two landfills (one flow-through with 10 and another recirculated leachate with a one-time point) leachate and studied the effect of PFAA-precursor transformation, leachate physical−chemical properties, leachate flow rates and meteorological variables. For the flow-through system, the authors observed an increase of ≥C8 PFCAs during March-mid April, which significantly correlated with 8:2 FTUCA, 8:2 FTCA, 10:2 FTUCA and 10:2 FTCAs indicating biotransformation of 8:2 and higher precursors (e.g., 8:2 PAP and FTOHs). In contrast, recirculated leachate generally contained an order of magnitude lower ΣPFAS concentrations, consisting entirely of PFAAs (∼83% PFCAs and ∼17% PFSAs on a molar basis). While the observed PFAS levels and profiles in the single sample of recirculated leachate might not be representative of leachate from this site over the long term, another explanation could be that recirculating leachate back into the landfill may have facilitated higher biotransformation and resulting in the absence of PFAA-precursors at this site. Other factors such as the volume of waste handled by each landfill and waste composition could also contribute to the observed variation (Benskin et al., 2012). No other studies have measured PFAA-precursors in recirculated landfill leachate; therefore, the observation of Benskin et al., (2012) could not be supported by other studies. Concentrations of PFBA, PFPeA, and PFHxA were reported to be strongly associated with increasing pH, electrical conductivity and decreasing 24 h precipitation in flow-through leachate (Benskin et al., 2012). This enhanced mobilization of PFASs with increasing pH is consistent with sorption studies of PFOS and PFOA to different adsorbents which indicated decreased sorption with increasing solution pH due to protonation of the adsorbent surface leading to fewer positive sites on the sorbent (Wang and Shih, 2011; Yu et al., 2009). This could imply that landfills undergoing methanogenesis (pH ≥ 7) are more conducive to PFASs leaching. Similarly, the observed seasonal variation of the macro-constituents (e.g., Cl-, Ca2+, Mg2+, SO42-) in leachate (Kulikowska and Klimiuk, 2008) likely contribute to the observed variability and patterns of PFASs concentration in leachates. A leaching study (Kim et al., 2015) using stain-guard treated carpets (24 h contact time) reported higher ∑PFASs partitioning in distilled water than in landfill leachate. The authors (Kim et al., 2015) attributed this to the presence of multivalent cations in the 21  leachate acting as bridges between anionic PFASs and negatively charged carpet surfaces and reduced desorption. In contrast to cations, anions (e.g., Cl-, SO42- or Cr2O72-) have been reported to compete with anionic PFASs for adsorption sites, (in boehmite, chitosan and resins) leading to increased solubility of anionic PFASs (Du et al., 2014). This shows that the effects of inorganic ions on PFAS leaching are complex and possibly concentration- and type-specific.  Allred et al., (2015) and Lang et al., (2016) studied the evolution of PFASs into leachate from MSW and carpets and clothing, respectively, using anaerobic landfill reactors. Both studies showed that PFASs are released through a combination of biotic and abiotic processes. However, at the end of the operation, the leachates from live bioreactors (producing methane) had on average 5 to 10 times higher ∑PFAS than the average for abiotic reactors, indicating that biological processes play an essential role (Lang et al., 2016; Allred et al., 2015). Following the onset of the methanogenic conditions, concentrations of known biotransformation intermediates of PFAA precursors, including methyl-perfluorobutane sulfonamide acetic acid, n:2 and n:3 FTCAs, increased steadily, with the 5:3 FTCAs becoming the single most concentrated PFAS observed in live reactors (Allred et al., 2015). In addition to precursor transformation, they also suggested that abiotic leaching, pH, type of substrate and sorption are likely to be influential factors determining PFASs profiles in leachate (Allred et al., 2015). Owing to the limited number of studies measuring PFAS compounds in landfill ambient air, the extent of release of semi-volatile precursors (e.g., FTOHs) and their volatile transformation products are unknown and need investigation. Li et al., (2015b) investigated adsorption and leaching of PFASs (10 PFCAs, 4 PFSAs, 1 FOSA and 3 FTUCAs) in sodium bentonite (common landfill liner material) under simulated field condition. The authors reported that PFASs from leachate did not bind substantially to the hydrophilic sodium bentonite, likely due to negative surface charges of PFAAs under environmental conditions. For the range of concentrations tested, PFASs did not affect the hydraulic conductivity of the bentonite, indicating that bentonite liners are not likely to be compromised by PFAS. However, the leaching cell test with sand/bentonite mixture showed partial retention of PFASs which decreased over time, indicating limited effectiveness of sodium bentonite liners in landfills containing PFASs (Li et al., 2015b). 22  2.6 Fate of PFASs in leachate treatment systems One of the most common methods for managing leachate is to send it to an off-site domestic WWTP. The fate and occurrence of PFASs in WWTPs have recently been reviewed elsewhere (Arvaniti and Stasinakis, 2015) and are outside the scope of this study. Other leachate management options include on-site pre-treatment followed by off-site discharge at a WWTP, and complete treatment and discharge on-site (Townsend et al., 2015). Leachate treatment options can be broadly categorized under physio-chemical treatment (e.g., coagulation-flocculation, chemical precipitation, membrane filtration, activated carbon adsorption, chemical oxidation) and biological treatment (e.g., activated sludge system, aerated lagoon, constructed wetlands) (Foo and Hameed, 2009; Renou et al., 2008). Several studies (Fuertes et al., 2017; Yan et al., 2015; Busch et al., 2010) have reported an overall increase in PFAA concentrations and a decrease in precursors (Yin et al., 2017) in the liquid phase following on-site biological leachate treatment processes. The extent of formation observed was analyte- and site-specific, ranging between 10 and 250% for individual PFAAs (Yan et al., 2015). In addition, removal of PFAAs was also reported in constructed wetlands through sorption by wetland soil and possibly plant uptake in the reed bed (Yin et al., 2017). Wet air oxidation process contacting with ozone to create OH-radicals to degrade contaminants also showed slightly higher (~5%) ∑PFAAs concentrations in the effluent leachate, but the increase was less than for a biological treatment (Busch et al., 2010). An adsorption technique using activated carbon has been reported to be somewhat useful (removal efficiency ranges between 70 and 99%) in removing PFAAs from leachate (Busch et al., 2010). High-pressure membrane filtration technologies such as reverse osmosis (RO) and nanofiltration (NF) removed >95% PFAAs directly from leachate (Busch et al., 2010) and biologically treated leachate (e.g., membrane bioreactor followed by RO or NF) (Yan et al., 2015). On the other hand, ultrafiltration (UF) integrated with membrane bioreactors showed little or no removal of PFAAs (Fuertes et al., 2017). Despite the success of high-pressure filtration systems, disposal of the PFAS-rich concentrate remains a challenging issue in need of careful consideration (Rahman et al., 2014).  2.7 Future research directions Compared to PFAAs, little information exists regarding their precursor compounds (e.g., fluorotelomer compounds) in the landfill. As reported by a recent study using total oxidizable 23  precursor assay, unknown precursors for C4 – C12 PFAAs contribute 10 – 97 mol% in leachate, which can account for additional 15%–43% mass loads (Wang et al., 2020). To avoid significant underestimation of the total PFAS released from landfills, PFAA-precursors (e.g., FTOH, FTS, etc.) and their transformations products should be included in future studies. As previous research (Phillips et al., 2007) reported 100-times smaller toxicity thresholds of FTCAs compared to PFCAs for freshwater microorganisms, PFAA-precursor concentrations in leachate would provide valuable information from water quality perspective. While two recent studies, using simulated anaerobic landfill reactors, have provided useful insights into the release of PFASs in the landfill, further research is needed to understand the transformation process fully. Studies need to include landfill gas to understand transformation pathways and the overall fate of PFASs, considering the semi-volatile nature of some precursor compounds and their transformation products. This would also enable more realistic assessments of the release of the PFASs to the environment with landfill gas and leachate. More research is also needed to evaluate the effectiveness of current containment practices (e.g., landfill liners) and how they can be improved to reduce PFAS emissions from landfills. While ubiquitous occurrences of various classes of PFASs have been studied and documented in eight developed countries, occurrence data is needed for landfills in developing countries.  The lack of regulations limiting the manufacture and use of C8-based PFASs and the lack of pollution abatement measures (e.g., leachate collection systems, lining materials) (Ismail and Manaf, 2013) mean landfills in developing countries could be a source of long-chain PFASs in the environment. This could undermine the regulatory initiatives in some parts of the world, due to the long-range transport and persistence of some PFASs (e.g., PFAAs). 2.8 Conclusions  This study presents a critical review of existing publications to identify temporal and geographical trends of PFASs occurrence and summarizes fate and transformation of PFASs in the landfills. Research over the past decade showed that PFAAs are routinely detected in landfill leachate. Short-chain (C4  ̶  C7) PFCAs are routinely detected in landfills from various countries, possibly indicating their greater mobility and reflecting the industrial shift towards shorter chain compounds. Despite its restricted use, PFOA remains one of the most frequently detected and abundant PFCAs in landfill leachate. If not appropriately managed, landfills could act as secondary 24  sources of PFOA in the environment. Recent studies also document the presence of PFAA-precursors and their transformation products in landfill leachate, at concentrations comparable to or higher than the most frequently detected PFAAs (e.g., PFBA, PFOA, PFOS). Landfill ambient air also contains elevated levels of PFASs, primarily semi-volatile precursor compounds (such as FTOHs), compared to upwind control sites. Therefore, landfills likely act as emission sources of atmospheric PFASs. The fate of PFASs inside landfills is controlled by a combination of biotic and abiotic processes, with biotransformation releasing most of the PFASs from landfilled waste to leachate. Biotransformation in simulated anaerobic reactors has been found to be closely related to the methanogenic phase. The methane yielding stage also results in higher pH (>7) of leachate, correlated with higher mobility of PFAAs.           25  Chapter 3: Aerobic Biotransformation of Fluorotelomer Compounds in Landfill Leachate-Sediment 3.1 Introduction Landfills are the primary disposal options for solid wastes (Kaza et al., 2018), including poly- and perfluoroalkyl substances (PFASs) containing end-of-life consumer products, construction wastes, biosolids, etc. Water percolating through landfilled waste (i.e., leachate) (Sui et al., 2017; Clarke et al., 2015) and landfill gas emission (Durmusoglu et al., 2010; Allen et al., 1997) are emerging as potential sources of organic contaminants, including PFASs, into the environment (Qi et al., 2018). In recent years, PFASs have gained notoriety due to their persistence, bioaccumulation potential and possible adverse effects on humans, animals and biota (Houde et al., 2011; Kato et al., 2011; Kannan et al., 2004). Fluorotelomer compounds (FTCs), belonging to PFASs, are used widely as water repellents on a wide range of finished products (e.g., textiles, apparel, carpet, fabrics), as oil and grease repellents on paper and packaging, and in other surface applications (e.g., paints, adhesives, polishes, metals, electronics) (Buck et al., 2011; Kissa, 2001). After disposal in landfills, consumer products containing FTCs undergo abiotic and biotic transformations (Washington et al., 2019; Washington et al., 2015b), releasing fluorotelomer monomers (e.g., fluorotelomer alcohols (FTOHs; CnF2n+1C2H4OH) and sulfonic acids (FTSs; CnF2n+1C2H4SO3H), as well as their transformation products (Lang et al., 2017; Allred et al., 2014; Benskin et al., 2012). Water-soluble monomers (e.g., FTSs) and their transformation products can then partition to landfill leachate and be released to aquatic environments or groundwater (Yang et al., 2014; Benskin et al., 2012). While neutral compounds like FTOHs can partition with landfill gas due to their low solubilities and relatively high vapour pressure (Weinberg et al., 2011), and subsequently enter the atmosphere (Figure 2.2 in Chapter 2). FTCs in the atmosphere can undergo long-range transport and further transformations resulting in the release of persistent perfluorocarboxylic acids (PFCAs) to remote environments (Filipovic et al., 2015; Yamazaki et al., 2016) (Figure 2.2 in Chapter 2). Previous studies have shown that FTCs can undergo biotransformation in various environmental media including activated sludge (Zhao et al., 2013a; Wang et al., 2012a; Wang et 26  al., 2011), river sediment (Zhang et al., 2017; Zhang et al., 2016; Zhao et al., 2013b) and soil (Liu et al., 2010c), leading to formation of saturated fluorotelomer acids (FTCAs), unsaturated fluorotelomer acids (FTUCAs) and persistent PFCAs. Environmental concerns regarding the persistence, ubiquitous occurrence and health effects of PFCAs are well documented (Houde et al., 2011; Kato et al., 2011). In May 2019, perfluorooctanoic acid (PFOA), a biotransformation product of 8:2 FTOH, was placed on the list of substances to be eliminated, as a persistent organic pollutant (POP) under the International Stockholm Convention on POPs (Stockholm Convention, 2019, Annex A). This has led manufacturers to shift toward short-chain (e.g., 6:2 FTS) alternatives of legacy FTC (e.g., 8:2 FTOH). Studies have shown endocrine disruptive activity of FTOHs in fish (Liu et al., 2009; 2010a; Ishibashi et al., 2008) and human cells (Rosenmai et al., 2013; Liu et al., 2010b; Maras et al., 2006), reproduction impairment in fish (Liu et al., 2010b) and rats (Mylchreest et al., 2005) and possible immunotoxicity in human (Kong et al., 2019). In addition, FTCAs are known to cause developmental toxicity to fish (Shi et al., 2017) and are orders of magnitude more toxic to aquatic organisms than their PFCA counterparts (Mitchell et al., 2011; Phillips et al., 2007). Previous studies (Zhang et al., 2016; Wang et al., 2005a, 2005b, 2009, 2011; Liu et al., 2007; Dinglasan et al., 2004) have investigated the fate and transformation of FTOHs and FTSs in various environmental conditions to assess the overall impact on the ecosystem and their contribution to PFCA exposure. Dinglasan et al., (2004) observed that 85 mol% of the spiked 8:2 FTOH was biotransformed by day 7 in a mixed microbial culture prepared from sediment and groundwater. Wang et al., (2005a, 2005b) studied biotransformation of 8:2 FTOH in activated sludge over periods of 28 (Wang et al., 2005a) and 90 days (d) (Wang et al., 2005b) and reported eight metabolites, including 8:2 fluorotelomer aldehyde (FTAL), 8:2 FTCA, 8:2 FTUCA, 7:3 FTCA, 7:3 FTUCA, 7:2 secondary fluorotelomer alcohol (sFTOH), 7:3 unsaturated amide and  PFOA.  Liu et al., (2007) observed first-order reaction rate of 8:2 FTOH biotransformation, ranging between 0.13 and 0.28 d-1 in clay loam soil, using three carrier-solvents (ethanol, octanol, 1, 4-dioxane), which may act as carbon sources. Similar to Wang et al., (2005a), the metabolites included a mixture of saturated and unsaturated fluorotelomer acids and PFOA. Perfluorohexanoic (PFHxA) and perfluoroheptanoic (PFHpA) acid were also observed, indicating that multiple -CF2 groups can be removed by microbial activity (Liu et al., 2007). Based on a study of aerobic 27  biotransformation of 8:2 FTOH in clay loam and sandy loam soils over 7 months, Wang et al., (2009) reported that 10 – 35 mol% of 8:2 FTOH was irreversibly bound to the soil and could not be solvent extracted. While the observed metabolites were similar to those of previous studies (Liu et al., 2007; Wang et al., 2005a, 2005b), the yields of PFOA (25%) and 7:3 FTCA (11%) were at least two orders of magnitude higher than other reported yields (Liu et al., 2007; Wang et al., 2005a; 2005b), suggesting that in some environmental matrices, significant PFOA exposure could result from 8:2 FTOH biotransformation (Liu and Mejia-Avendaño, 2013). Relatively limited data and conflicting results are available for the biotransformation of FTSs (Field and Seow, 2017). Aerobic biotransformation of 6:2 FTS resulted in the formation of similar products consisting of mixtures C6 or shorter chain PFCAs and FTCAs in activated sludge and river sediments. However, wide variations in the biotransformation rate of 6:2 FTS and product formation have been reported in various environmental media (e.g., t1/2 <5 d in river sediment, >2 years in activated sludge) (Zhang et al., 2016; Wang et al., 2011).  Based on historical production volume of FTCs (estimated 11,000 to 14,000 t/year) (Lassen et al., 2013), the magnitude of landfill emissions could amount to several-fold (Washington et al., 2019; Washington and Jenkins, 2015) the currently estimated global oceanic load of C4 – C14 PFCAs (~10,000 t) (Wang et al., 2014). While two studies (Lang et al., 2016; Allred et al., 2015) have investigated the release of PFASs from municipal solid waste using anaerobic landfill reactors, aerobic biotransformation of FTCs in landfill leachate has not been studied to date. Previous studies (Wang et al., 2005a; 2009) have used simple solutes (e.g., deionized water), which are likely, not representative of complex environmental matrices (e.g., landfill leachate). In general, landfill leachate is characterized by high total ammonium nitrogen (hundreds to several thousand of mg/L), biodegradable and refractory organic matter (tens to several thousand mg/L), high alkalinity (hundreds to several thousand mg CaCO3/L), xenobiotic organic compounds and heavy metals (Ren et al., 2017; Gao et al., 2015; Renou et al., 2008). Furthermore, the microbial community in landfill leachate is expected to differ from those in surface soil, river sediment, and activated sludge from municipal WWTPs, which can affect the rate and metabolite yield of FTCs. The objective of the current investigation was to study the aerobic biotransformation of FTCs in the leachate-sediment system to determine the known metabolite yields. Landfill leachate was used for microcosm preparation in this study. The effect of long-chain (8:2 FTOH) versus short-chain 28  (6:2 FTS) FTC in PFCA production was also investigated. The results of the current investigation are compared with findings from previous studies of FTCs biotransformation in other environmental media.  3.2 Materials and methods  Standards and reagents The analytes of interest, comprising of the parent FTCs and their known biotransformation products, are listed in B.1 of Appendix B with their acronyms, CAS numbers and suppliers. The monitored compounds included perfluorobutanoic acid (PFBA), perfluoropentanoic acid (PFPeA), PFHxA, PFHpA, PFOA, perfluorononanoic acid (PFNA), 8:2 FTCA, 7:2 FTCA, 8:2 FTUCA, 7:2 FTUCA, 7:3 FTCA, 5:3 FTCA, 8:2 FTOH, 6:2 FTOH, secondary fluorotelomer alcohols (7:2 sFTOH and 5:2 sFTOH) and 6:2 FTS. HPLC grade methanol, acetonitrile, ammonium acetate and acetic acid were purchased from Fisher Scientific, Canada. For all purposes, ultrapure deionized water (DI) from a Millipore® system was used. Oasis weak anion exchange (WAX) (6 cc/30 µm) and Sep-Pak C18 Plus Short Cartridge were purchased from Waters (MA, USA).   Landfill sediment and leachate collection Sediment was collected in December 2017, from a leachate collection ditch in a municipal landfill (location kept confidential at the request of the operator) to set up the biotransformation microcosms. The sampling equipment and high-density polyethylene (HDPE) sample collection bottles were washed thoroughly, rinsed with methanol and air-dried prior to sampling. Using a steel scoop, the top 1 cm of the sediment layer was collected to ensure the presence of aerobic microbial communities (Marchant et al., 2017). Landfill leachate was collected in 3L HDPE bottles from the surface of the ditch, at the same location as the sediment collection. The sampling bottles were half-filled with either sediment or leachate to ensure aerobic conditions during transportation. Sediment and leachate samples were transported to the lab within an hour of collection. Collected sediment was wet-sieved via a 2 mm sieve and stored with 40 – 50 mm of leachate on top at 4°C. Within 2 d of collection, initial physical and chemical characteristics of the leachate and sediment were determined (section B.2 of Appendix B), followed by the set-up of the microcosms.  29   Microcosm setup and sampling A closed-system microcosm set-up using Wheaton glass serum bottles (125 mL), fitted with aluminum crimp-sealed natural rubber stoppers (Figure B.1 of Appendix B), was adopted from Liu et al., (2010c) and Yu et al., (2016) to minimize loss of volatile transformation products. The experimental conditions and appropriate controls are provided in Table 3.1. In each bottle, 15 g of wet sediment, 10 mL of leachate and 5 mL of mineral media were added. The mineral medium solution was prepared by adding 85 mg KH2PO4, 218 mg K2HPO4, 334 mg Na2HPO4·2H2O, 5 mg NH4Cl, 36.4 mg CaCl2·2H2O, 22.5 mg MgSO4·7H2O, and 0.25 mg FeCl3·6H2O in one liter DI (OECD, 1992). The chemicals used for mineral media preparation and their suppliers are listed in B.3 of Appendix B. Sterile control microcosms were prepared by three cycles of autoclaving (at 121°C for 1 h) with intermittent incubation (20°C for 24 h) between cycles. In addition, 0.75 g/L of NaN3 was added to each sterile bottle to prevent microbial growth in the sediment mixture during incubation. Following a week of acclimation at 20°C, five treatments were applied to the microcosms as outlined in Table 3.1. All sample bottles were continuously agitated at 150 rpm on a temperature-controlled orbital shaker at 20°C in the dark (Innova 4200, New Brunswick Incubator Shaker). A total of nine sampling events occurred (at 0, 1, 3, 7, 14, 30, 45, 60, 90 d) over the three-month experimental period, with two bottles of each treatment condition sampled for each sampling event. The headspace concentration of O2 was measured using a Quantek oxygen analyzer 905 (Grafton, MA, USA) in the sterile and live-controls. All remaining bottles were reaerated when the O2 level approached the recommended minimum level of 10% oxygen saturation (i.e., 2% headspace O2 content) (Mejia-Avendaño et al., 2016; Liu and Mejia-Avendaño 2013) (Figure 3.1a). On each sampling day, 15 mL of the live-control microcosms (treatment 5 in Table 3.1) were collected to monitor the pH. In addition, total organic carbon (TOC), ammonia, nitrate and nitrite were determined using the dissolved phase of the live-controls (see B.2 of Appendix B). Given the high ammonia content of landfill leachate and the use of biological nitrification as a leachate treatment method, nitrogen redox was monitored for possible nitrification occurring in the live microcosms. 30  During sample collection, the headspace gas of live-spiked and sterile microcosms (treatments 1 through 4 in Table 3.1) were sampled to capture volatile compounds by flushing with ~1 L air through two C18 cartridges. The cartridges were then eluted with 5 mL acetonitrile, and the eluents were stored at -20 °C until PFAS analysis using GC-MS/MS. Table 3.1 Experimental conditions of biotransformation microcosms containing sediment, landfill leachate and mineral media Treatment Microcosm Parent compound spike Inhibitor 1 Live-spiked 8:2 FTOHa - 2 Sterile-control 8:2 FTOH NaN3 3 Live-spiked 6:2 FTS - 4 Sterile-control 6:2 FTS NaN3 5 Live-control None - a8:2 FTOH and 6:2 FTS spiking stock concentrations were 50 mg/L and 250 mg/L, respectively, prepared in methanol.   Microcosm extraction and PFAS analysis For each sampling event, duplicate microcosms were subjected to two sequential extractions. The solvent extraction method was modified from that of Liu et al., (2010c). Before the first extraction, the crimp seal of the bottle was opened, and the cap was pushed inside to be extracted at the same time. 30 mL of acetonitrile was added to each bottle and the cap was resealed. The bottles were placed on a horizontal shaker (180 rpm) at 40°C for 3 d, then centrifuged at 1000 rpm (~160 g) for 20 min to collect the supernatants. For the second extraction, 27 mL acetonitrile and 3 mL of 250 mM NaOH were added to the bottles. The bottles were next shaken at 40°C overnight, centrifuged at 1000 rpm (~160 g) for 20 min, and the collected extracts were stored at -20 °C for PFAS analysis. Microcosm extracts from all five treatment conditions were cleaned-up using solid-phase extraction (SPE) with Oasis WAX® following a modified method (USEPA, 2011). Microcosm extract (25 mL) was spiked with 70 ng of each M8PFOA and M2_8:2 FTUCA, diluted to 100 mL with ultrapure water and adjusted to a pH of 6.5±0.5 with 1% (v/v) acetic acid in reagent water or 0.3% (v/v) aqueous ammonium hydroxide. The extract was then cleaned-up using a weak anion 31  exchange (WAX) solid-phase extraction (SPE) cartridge. Each cartridge was conditioned with 0.3% NH4OH in methanol and 0.1 M formic acid in DI, followed by equilibration with DI. After loading the sample drop-wise onto the cartridge (~5 mL/min), the cartridge was washed with 20% MeOH in 80% 0.1 M formic acid in DI, followed by 0.3% (v/v) NH4OH in DI. The cartridge was then dried by being subjected to a vacuum for 5 min. Finally, the cartridge was eluted into a 15-mL clean glass centrifuge tube with 4 mL 0.3% NH4OH (v/v) in MeOH. The cleaned-up extract was then dried under a gentle nitrogen stream, reconstituted with 980 µL of injection solvent (80:20/water: acetonitrile) and spiked with injection internal standard solution (containing 30 ng of M3PFBA and M5PFHxA) prior to instrumental analysis.  Headspace extracts from all spiked conditions and microcosm extracts from 8:2 FTOH spiked live and sterile treatments (outlined in experimental conditions in Table 3.1) were also analyzed using GC-MS/MS to quantify volatile parent and transformation products. Microcosm extracts filtered through 0.45 µm filter and headspace extracts were spiked with 70 ng of (M+4) 8:2 FTOH as internal injection standards before analysis using GC-MS/MS.  Instrumental analysis of PFAS  An Agilent 1200 series HPLC system (Agilent Technologies, CA, USA) was used to separate the PFCAs using 4 µL of sample injection onto a Waters Xterra MS C18 column (100×2.1 mm, 3.5 µm particle size; Waters Corporation, Milford, MA), preceded by an Xterra MS C18 VanGuard cartridge (30×2.1 mm, 3.5 µm particle size) from the same manufacturer. Both columns were maintained at 50°C, and the mobile phase flow was set at 0.7 mL/min. The mobile phase consisting of (A) water with 10 mM ammonium acetate and 10 mM acetic acid, and (B) acetonitrile (Gradient and detailed instrumental conditions are specified in B.4 of Appendix B). Mass spectrometric analysis was performed using an Agilent 6410 Triple Quad mass spectrometer (Agilent Technologies, CA, USA) in negative electrospray ionization and multiple reaction monitoring (MRM) modes. MS parameters of each analyte (see B.4 of Appendix B) were individually optimized manually during syringe pump infusion.  The leachate-sediment microcosm and headspace cartridge extracts were also analyzed on an Agilent Technologies (Palo Alto, CA) 7890B gas chromatograph, interfaced with a 5977B mass-32  selective detector operated in the positive chemical ionization (PCI) mode with methane reagent gas for quantitative analyses. Analyte separation and quantification were performed on an Agilent CP-Sil 8 CB column (30 m x 0.25 mm I.D., 0.25 µm film thickness). Samples (1 µL) were injected in the pulsed-splitless mode at ~276 kPa (40 psi) for 0.015 min. The heating profile of the column and MS parameters are detailed in B.5 of Appendix B.  Experimental and analytical quality control and quality assurance High-density polyethylene (HDPE) and polypropylene (PP) were used throughout the sample preparation and storage stage. Polytetrafluoroethylene (PTFE) containing material was eliminated to avoid possible contamination of FTCs. All experiments were performed in duplicates. During sample preparation, procedural blanks were included to account for any PFAS contamination. Any LC-MS/MS or GC-MS/MS run was accompanied with quality control check standards and a solvent blank every 10 – 15 samples to monitor the absolute analyte areas, chromatographic retention times and background contamination from the instrument. 3.3 Results and discussion  Physical and chemical characteristics of the landfill leachate The initial characteristics of the leachate used in this study are compared with the ranges of values reported in previous studies of mature landfills (>10 years) in Table 3.2. Since the landfill leachate was collected from a leachate collection ditch, it likely represents a mixture of leachate from various stages of waste decomposition. However, as shown in Table 3.2, the basic pH (~8) of the leachate indicated that the landfill was primarily in the methanogenic phase, consistent with the fact that the municipal landfill which provided the leachate for this study, has operated for more than 50 years and can be considered mature (Renou et al., 2008). Young leachate produced from refuse undergoing an early stage of stabilization (e.g., acid phase) is also expected to reflect the composition of methanogenic leachate if the young leachate percolates through well-decomposed refuse (Kjeldsen et al., 2002). Any surface runoff from precipitation at the landfill site is also treated as leachate and collected in the same ditch. Since the landfill site received >150 mm rainfall in the 30 days prior to the sampling day, the collected landfill leachate can be considered dilute compared to dry season leachate at this site or other landfills located in the arid climate. Dilution 33  of leachate by precipitation also supports the observation that the measured organic content, as well as the anion and cation concentrations of the leachate, were closer to the lower bound of literature values than the higher bound, as shown in Table 3.2.  Table 3.2 Comparison of leachate characteristics with literature values Leachate Characteristics This study (n=3) Literature values for mature (>10 years) leachate References pH 7.9 ± 0.1 7.5 -9 Kjeldsen et al., 2002 Alkalinity (expressed as mg CaCO3/L) 710 ± 1% 600 – 17,000 Robinson, 2007; Kjeldsen et al., 2002 Total suspended solids (mg/L) 86 ± 6% N/Aa - Total organic carbon (mg/L) 304 ± 8% 100 - 5000 Robinson, 2007 Ammonia (mg/L) 84 ± 4% 100 - 2500 Robinson, 2007; Kjeldsen et al., 2002 Metalsb (mg/L)    Calcium 93 ± 11 60 - 600 Robinson, 2007; Kjeldsen et al., 2002 Iron 5.2 ± 0.4 3 - 280 Kjeldsen et al., 2002 Magnesium 23.6 ± 0.7 40 - 350 Kjeldsen et al., 2002 Anions (mg/L)    Fluoride 1 ± 50% N/A - Chloride 188 ± 0.5% 200 - 3000 Kjeldsen et al., 2002 Sulfate 31.3 ± 1.6% 10 - 420 Kjeldsen et al., 2002 Sediment Characteristics    Electrical Conductivity (mS/m) 6.8 N/A  Moisture Content (%) (wet wt. basis) 74 ± 0.2%   Organic Content (%) (dry wt. basis) 1.8 ± 0.1%   aN/A: not available; bmetals higher than the detection limit (0.5 mg/L) are shown    Microcosm operating conditions of live controls The variation of headspace oxygen content, pH, TOC, ammonia and nitrate for the live-control (no spike) microcosms over 90 d of operation are shown in Figure 3.1. The initial headspace oxygen content (16%) decreased throughout the experiment, indicating aerobic respiration. The headspaces were reaerated by flushing with air at 30 d (Figure 3.1a) to maintain aerobic conditions throughout the experimental period. The observed minimum headspace oxygen content (6%) was higher than the recommended minimum of 10% oxygen saturation (i.e., 2% oxygen content; v/v) (King et al., 1997) used in biotransformation studies (Mejia-Avendaño et al., 2016). The live-control microcosms pH was ~8 on day 0, dropping to ~7 by day 14 and then remaining nearly constant throughout the rest of the experiment (Figure 3.1b). The initial drop in pH and faster 34  oxygen depletion within the first 30 days coincided with observed nitrification (Figure 3.1c) in live-controls. Significant alkalinity was lost to neutralize the hydrogen ions released during ammonia oxidation (Equation 3.1), the first step of nitrification (Equations 3.1 and 3.2).  2NH4+ + 3O2 → 2NO2-+ 4H+ + 2H2O                                                                         (3.1) 2NO2-+ O2→2NO3-                                                                                                                                                         (3.2)  Figure 3.1 Time variation of a) headspace oxygen content, b) pH, c) nitrogenous compounds, d) total organic carbon (TOC) in the dissolved phase of the live controls (without any spiking). Absolute range of duplicate samples (n=2) is shown by error bars.  Nitrifiers also use inorganic carbon (e.g., bicarbonate) as a carbon source. The high alkalinity (i.e., buffering capacity) of landfill leachate (Table 3.2) favoured the growth of nitrifiers and did not allow the pH to drop excessively (Figure 3.1b). It is known that nitrification can occur by autotrophic and heterotrophic bacterial activity (Watson et al., 1981). Autotrophic bacteria are thought to be primarily responsible for aerobic nitrification in the natural environment, as 35  heterotrophic nitrifiers have significantly slower kinetics (Braker and Conrad, 2011; Ward B., 2011; Tortoso and Hutchinson, 1990). Recent studies have reported higher PFCA formation in the presence of increased ammonia-oxidizing activity (Yu et al., 2018a and 2018b). Therefore, the activity of nitrifiers in landfill leachate is likely to affect the biotransformation of 8:2 FTOH and 6:2 FTS. Nitrifying bacteria are obligate aerobes; the near-complete nitrification in this study suggest that aerobic condition prevailed in the microcosms throughout the experiments. The initial TOC of the dissolved phase (~200 mg/L) decreased slowly throughout the experiments and reached an average removal of 35% by day 90 (Figure 3.1d). The incomplete removal of TOC is likely due to the presence of recalcitrant organics in the leachate (Kjeldsen et al., 2002). Overall, oxidation of leachate organics indicates the presence of aerobic heterotrophic bacteria in the live microcosms.   FTCs and PFCAs in background landfill leachate and sediment Among the known metabolites of parent compounds, PFPeA, PFHxA, PFHpA, PFOA, 7:3- and 5:3 FTCAs were observed in the sterilized and live matrix controls at comparable levels, with little variation throughout the biotransformation experiment (Figure 3.2). The n:3 FTCAs observed in live-control microcosms (Figure 3.2b) are not known to be manufactured and are almost exclusively formed through biotransformation of FTCs (e.g., 8:2 FTOH, 6:2 FTOH, 6:2 FTS etc.). Therefore, the presence of n:3 FTCAs in live-control microcosms (Figure 3.2b) at day 0 indicated that FTC biotransformation was likely occurring in the leachate collection ditch from which the sediment and landfill leachate were sampled. At 90 d, the sum of the quantified products in live-controls was equivalent to ~1 mol% of the 8:2 FTOH and ~0.2 mol% of the 6:2 FTS spiked at day 0, significantly less than observed in the live-spiked microcosms, as discussed below in section 3.3.4. 36   Figure 3.2 Background concentrations of a) perfluorocarboxylic acids (PFCAs) in live-controls (no 8:2 FTOH/6:2FTS spike); b) fluorotelomer acids live-controls; c) PFCAs in sterile-controls; d) fluorotelomer acids sterile-controls throughout the experimental period. Absolute ranges of the duplicate samples (n = 2) are expressed by error bars. Error bars smaller than the symbol height are not visible.  Biotransformation of spiked parent compounds in landfill leachate and sediment The biotransformation products of 8:2 FTOH and 6:2 FTS are shown in Figure 3.3. The biotransformation rates, products and yields observed in this study are compared with those from previous studies in Tables 3.3 and 3.4 for 8:2 FTOH and 6:2 FTS, respectively. Mass balances of the spiked 8:2 FTOH and 6:2 FTS are presented in Figure 3.4.  Biotransformation products of spiked 8:2 FTOH and 6:2 FTS Production of known biotransformation products (Figures 3.3a and b), along with the decreasing trend of 8:2 FTOH in live-spiked microcosms (Figure 3.4a), suggest that 8:2 FTOH was undergoing aerobic biotransformation in the live landfill leachate-sediment microcosms. Six known metabolites of 8:2 FTOH were quantified (Figure 3.3a and b) with molar yield > 9% of the 37  initially spiked 8:2 FTOH at 90 d. The observed product yield was slightly less than previously reported yields of 8:2 FTOH products in soil and activated sludge, which ranged between 10 to 40% for experimental durations of 30 to 210 d (Wang et al., 2005a, 2005b; 2009) (Table 3.3). 8:2 FTOH gradually degraded into four main products PFHxA, PFHpA, PFOA and 7:3 FTCA (sum >7 mol% yield after 90 d), with PFOA being the most abundant product (~2.8%). While most previous 8:2 FTOH biotransformation studies did not observe PFHpA in pure culture, mixed culture and soil (Table 3.3), Yu et al. (2016), using activated sludge from an industrial WWTP, reported generation of PFHpA (0.21 mol% after one day) from 8:2 FTOH under aerobic conditions. Transient metabolites included 7:2s FTOH and 8:2 FTUCA.  Although 7:3 FTUCA, a known precursor of 7:3 FTCA, was detected in LC-MS/MS (Figure B.2 of Appendix B), it was not quantified due to the lack of an authentic standard. 8:2 FTCA, a precursor of 8:2 FTUCA was not observed in any of the live-spiked samples, likely due to its rapid transformation to 8:2 FTUCA by microbes (Wang et al., 2009). Also, defluorination of 8:2 FTCA to form 8:2 FTUCA could occur abiotically during sample extraction or storage in the presence of a base (Liu and Mejia-Avendaño, 2013; Wellington Laboratories, 2012). No peaks were observed for relatively newly identified products 3-hydroxy-7:3 saturated fluorotelomer carboxylate (3-OH-7:3 FTCA), 2H-polyfluorooctanoic acid (2H-PFOA) and unsaturated perfluorooctanoic acid (uPFOA) (Washington et al., 2015c). 38  Table 3.3 Comparison of experimental conditions and major outcomes of previous and current FTOH aerobic biotransformation studies  ahalf-life was calculated based on loss of parent compound (total 9 data points as shown in Figure B.3 of Appendix B); bN.A: not available; Parent compound Degradation media  Duration (d) Spiking concentration Stable products (yield in mol%) Half-life (d) Reference 8:2 FTOH  Landfill leachate and sediment 90  ~0.5 µmol/L PFOA (2.8 %), 7:3 FTCA (1.5%), PFHxA (0.5%), PFHpA (0.5%), 8:2 FTUCA (1.8%) >365a (R2 = 0.77) This study Ethanol degrading mixed microbial culture 80 750 µg/L 8:2 FTCA, 8:2 FTUCA (prominent), PFOA  N.Ab Dinglasan et al., 2004 Soil and pure culture of soil microbes Up to 67 100 µg of 8:2 FTOH/g soil 7:2 sFTOH, PFOA, PFHxA, 8:2 FTCA, 8:2 FTUCA, 7:3 FTCA, and 7:3 FTUCA N.Ab Liu et al., 2007 Industrial WWTP sludge 1 50 µg/L PFHxA (0.18%), PFHpA (0.21%), PFOA (2.24%) <1 Yu et al., 2016 14C labelled 8:2 FTOH Soil (14C labelled and native 8:2 FTOH) 210 20 µg/g dry soil PFOA (25%), 7:3 FTCA (11%), 2H-PFOA <7  Wang et al., 2009 Activated sludge 120 986 – 1049 µg/L  7:2 sFTOH, 7:3 FTUCA, 7:3 U amide, 8:2 FTAL, 8:2 FTCA, 8:2 FTUCA, 7:3 FTCA, PFOA, PFHxA - Wang et al., 2005b Diluted WWTP sludge 28  306 (±9) µg/L  8:2 FTCA (27%); 8:2 FTUCA (6%); PFOA (2.1%) - Wang et al., 2005a 4:2, 6:2 and 8:2 FTOH P. oleovorans and P. butanovora bacterial culture 28 ~0.8 µmol/L PFOA (7.9%), 7:3 FTUCA (2.9%), 7:3 FTCA (1.8%), PFHxA (0.62%), 8:2 FTUCA (~20%), 7:2 sFTOH (~18%) N. A. Kim et al., 2012 6:2 FTOH  Mixed bacterial culture and aerobic soil  90 d (bacterial culture) 180 d (soil) 2.8 – 20 µg/mL (bacterial culture) 2.9 µg/g (soil) 6:2 (6%) FTCA, 6-2 (23%) FTUCA, 5-2 sFTOH (16%), 5-3 acid (6-15%), PFHxA (5-8%) -mass balance in PFPeA (30%), PFBA (2%) <2  Liu et al., 2010c River sediment 100 - 5:3 FTCA (22.4%), PFPeA (10.4%), PFHxA (8.4%), PFBA (1.5%), 4:3 FTCA (2.3%) <2 Zhao et al., 2013a Activated sludge 60 1.57 – 3.13 mg/L PFHxA (11%), 5:3 FTCA (14%), PFPeA (4%) <2 Zhao et al., 2013b River sediment 28 5 – 15 mg/L 5:2 sFTOH (28%) 5:3 FTCA, (9.6%), PFHxA (11%) in high dose; 5:2 sFTOH (73%), 5:3 FTCA (23%), PFHxA (26%) in low dose <3 Zhang et al., 2017 [1,2-14C] 6:2 FTOH Soil 84 2.9 µg/g 5:2 sFTOH, 5:3 FTCA (12%), PFHxA (4.5%), PFPeA (4.2%) 1.3  Liu et al., 2010b 39  Table 3.4 Comparison of experimental conditions and major outcomes of previous and current 6:2 FTS biotransformation studies under aerobic condition  ahalf-life was calculated based on loss of parent compound (total 9 data points); bN.A: not available    Degradation media  Duration (day) Spiking concentration Stable products (yield in mol%) Half-life (day) Reference Landfill leachate and sediment 90 ~3 µmol/L PFBA (0.6%), PFPeA (5.6%), PFHxA (3.1%), 5:3 FTCA (0.2%), 5:2 sFTOH (0.6 %) ~86a  (coefficient of determination R2 = 0.86) This study Activated sludge 90 2.2 mg/L 5:3 FTCA (0.12%), PFBA (0.14%), PFPeA (1.5%), and PFHxA (1.1%) ~730 Wang et al., 2011 Gordonia sp. strain NB4-1Y 5  N.A.b 5:3 FTCA, 6:2 FTCA, 6:2 FTUCA and 5:3 FTUCA Not reported Van Hamme et al., 2013 Activated sludge Upto 40 weeks 500 mg/L No observable biotransformation - Ochoa-Herrera et al., 2016 River sediment 90 2.8 nmol/L 5:3 FTCA (16%), PFPeA (21%) and PFHxA (20%) <5  Zhang et al., 2016 Gordonia sp. strain NB4-1Y  7 N.A. 5:2 FT ketone (43.9%), 5:2 sFTOH (8.97%), and 6:2 FTOH (4.14 %) <7 Shaw et al., 2019  Figure 3.3 Formation of a) fluorotelomer compounds from 8:2 FTOH, b) perfluorocarboxylic acids (PFCAs) from 8:2 FTOH, c) fluorotelomer compounds from 6:2 FTS and d) short-chain PFCAs from 6:2 FTS in landfill leachate over the experimental period of 90 days. Absolute ranges of the duplicate samples (n = 2) are expressed by error bars. Error bars smaller than the symbol height are not visible. The average molar yield of the six quantified products (Figures 3.3c and d) of 6:2 FTS was ~10.8% at the end of 90 d, slightly higher than previously reported (Wang et al., 2011) yields in activated sludge (6.3% at 90 d). However, the observed yield in this study is lower than that in river sediment (>57% at 90 d) (Zhang et al., 2016) and in pure culture of Gordonia sp. strain NB4-1Y (Shaw et al., 2019; Van Hamme et al., 2013), that can metabolize 6:2 FTS as a sole sulfur source in sulfur-limiting conditions (>57% after 7 d) (Table 3.4). PFHxA, PFPeA, PFBA, 5:3 FTCA and 5:2 sFTOH were the terminal products in our study, with PFPeA being the most abundant product on day 90 (5.6%), followed by PFHxA (3.1%). Previous studies (Zhang et al., 2016; Wang et al., 2011) using mixed microbial communities also reported PFPeA and PFHxA to 41  be the primary end products, suggesting that aerobic biotransformation of 6:2 FTS is a source of short-chain PFCAs to the environment.  Comparison of the biotransformation products quantified here under aerobic conditions with those of the reported PFAS released from anaerobic landfill reactors (Lang et al., 2016; Allred et al., 2015) indicates that PFCAs and FTCAs are primary contributors, suggesting the formation of similar products. The authors (Lang et al., 2016; Allred et al., 2015) also reported the persistence of FTS compounds (e.g., 6:2 FTS) in the leachate, which will likely undergo biotransformation under aerobic condition following collection and removal of the leachate.  Biotransformation rates and products yield of 8:2 FTOH and 6:2 FTS 8:2 FTOH biodegraded gradually in aerobic live samples with ~30 mol% remaining in sediment-leachate solution after 90 d.  Wang et al., (2009) reported rapid biotransformation (>90% removal) of 8:2 FTOH within in aerobic soil, with early transient metabolites like 8:2 FTUCA reaching peak concentrations within 2 – 3 d, whereas stable metabolites (e.g., PFOA) achieved steady-state concentrations after 7 – 14 d. In our study, as shown in Figure 3.3a, 8:2 FTUCA reached a peak concentration after 60 d (~1.4 mol%), whereas PFOA needed 90 d to reach a steady-state concentration, with average yields of ~2.8 mol% (Figure 3.3b). Generally, 8:2 FTOH half-lives in the range of > 7 d have been reported in aerobic soils (Table 3.3). However, Washington et al. (2015b) found that 8:2FTOH had a half-life of ~210 d in saturated Appling Soil, which is closer to the value calculated in this study (>1 year). The slower biotransformation could be attributed to the cometabolic nature of FTOH biotransformation (Lewis et al., 2016; Kim et al., 2012). Cometabolic biotransformation allows the transformation of a non-growth substrate (e.g., FTOH) in the presence of a growth substrate (e.g., organic carbon, ammonia) (Dalton and Stirling 1982), making the process depend on the concentration and nature of the growth substrates (Fischer and Majewsky, 2014). Although the presence of growth substrates is a requirement for cometabolic biotransformation, orders of magnitude higher concentrations of the growth substrates may reduce the transformation of the micropollutant through competitive enzyme inhibition (Fischer and Majewsky, 2014; Plósz et al., 2009; Alvarez-Cohen and Speitel, 2001). The high dissolved organic carbon (Figure 3.1d) in the live-controls indicates additional carbon sources, which could retard the biotransformation of 8:2 FTOH (Lewis et al., 2016). 7:2 sFTOH, a previously reported (Liu et 42  al., 2007; Wang et al., 2009) major transient metabolite of 8:2 FTOH, formed during the initial stage, was quickly converted to downstream products (e.g., PFOA). In this study, 7:2 sFTOH remained below the quantification limit in the headspace until 23 d and accounted for <0.2 mol% after 60 d (Figure 3.3a), suggesting the slow transformation of 8:2 FTOH to 8:2 FTUCA  Another transient metabolite, 7:3 FTUCA, was observed in the sediment-leachate extracts but was not quantified due to the lack of an authentic standard (Figure B.1). The metabolite reached a peak, around 20 – 30 d and then decreased rapidly. Three stable products, PFHpA, PFHxA and 7:3 FTCA, reached steady-state levels after 23 d, with average yields of ~0.3, 0.4 and 0.6 mol%, respectively (Figure 3.3b).  Biotransformation of 6:2 FTS in landfill leachate was incomplete (>50% remaining) after 90 d, similar to what has been reported in activated sludge (Wang et al., 2011). Transient metabolites like 6:2 FTUCA reached maximum levels at around 30 – 45 d and then fell below the detection limits, resulting in an increase in downstream product (e.g., 5:2 sFTOH) concentrations (Figure 3.3c). More abundant end products (e.g., PFHxA and PFPeA) were still increasing after 90 d (Figure 3.3d). The lack of significant accumulation of the transformation intermediates (Figure 3.3c) suggests that the initial desulfonation was likely the rate-limiting step in 6:2 FTS biotransformation. A previous study (Shaw et al., 2019) with pure microbial culture showed that limiting sulfur can accelerate 6:2 FTS biotransformation. Therefore, alternate sulfur availability in landfill leachate (sulfate concentration ~0.32 mM, Table 3.2) might have contributed to the observed slower kinetics of 6:2 FTS biotransformation. In addition, the cometabolic nature of 6:2 FTS biotransformation suggests that the presence of large numbers of xenobiotic organic compounds in the landfill leachate (Clarke et al., 2015; Eggen et al., 2010) increases competition for the active sites in enzymes, resulting in slower biotransformation. 3.3.6.1 Comparison of 8:2 FTOH and 6:2 FTS product yields 6:2 FTS yielded products more quickly than 8:2 FTOH, as the time to reach a peak for 6:2 FTUCA was <45 d (Figure 3.3c), compared to ~60 d for 8:2 FTUCA (Figure 3.3a). In addition, higher accumulation of PFCAs was observed for 6:2 FTS compared to 8:2 FTOH (Figures 3.3b and d). Longer fluorinated chain and neutral functional groups in 8:2 FTOH could result in lower 43  solubility, leading to lower bioavailability compared with 6:2 FTS, resulting in lower biotransformation potential (Liu and Mejia-Avendaño, 2013). Similarly, 5:3 FTCA, a major product of 6:2 FTS in river sediment (Zhang et al., 2016) and activated sludge (Wang et al., 2011), did not show accumulation in leachate (Figure 3c), suggesting that it was likely undergoing further biotransformation to produce smaller chain PFCAs (e.g., PFBA, PFPeA) (Liu and Mejia-Avendaño, 2013; Wang et al., 2012a). 5:3 FTCA, also reported to be a dominant compound in leachate in U.S. landfills under anaerobic conditions (Lang et al., 2017), would likely undergo further biotransformation when the leachate is subjected to aerobic conditions. In contrast, during the corresponding periods, products of 8:2 FTOH, (i.e., 7:3 FTCA shown in Figure 3.3a), were accumulating in the leachate-sediment system and accounted for >1 mol% of the initially spiked 8:2 FTOH after 90 d. Higher biotransformation potential of the 6:2 FTS and its transformation intermediates imply that the shift towards short-chain FTC likely results in higher concentrations of short-chain PFCAs in landfill leachate, and possibly also in wastewater.  Mass balances of 8:2 FTOH and 6:2 FTS The overall mass balances representing the sum of recovered 8:2 FTOH and 6:2 FTS from spiked microcosms are shown in Figures 3.4a, b and c, respectively. The observed recovery of 8:2 FTOH (81 – 100%) from sterile control samples (Figure 3.4b), was in a range similar to that reported for FTOHs recovery from sediment (82 – 122%) (Zhao et al., 2013b) and soil (87 – 113%) (Liu et al., 2010c). The recovery of 81 – 100% of the spiked 8:2 FTOH demonstrates the integrity of the microcosm set-up and the efficiency of our extraction methods. However, the mass of 8:2 FTOH recovered from live-spiked microcosms showed a decreasing trend, reaching ~65% after 90 d (Figure 3.4a). By day 30, 8:2 FTOH recovered from headspace accounted for ~17% of dosed 8:2 FTOH, which increased slightly following reaeration at day 30 and averaged around 24% at 90 d. For the sterile microcosms, 19 % of the spiked 8:2 FTOH partitioned to the headspace of the sterile microcosms by 14 d, and it varied between 19 and 21% after that (Figure 3.4b). In general, previous studies (Wang et al., 2005b, 2009; Liu et al., 2007) reported higher partitioning of FTOHs in the headspace of sterile microcosm compared with live microcosms. However, previous studies have also observed rapid biotransformation (half-life <7 d) of spiked 8:2 FTOH, limiting availability of 8:2 FTOH to headspace from soil or activated sludge in live microcosms (Liu et al., 2007; Wang 44  et al., 2005a, 2005b, 2009). It is likely that the slower biotransformation observed by us in landfill leachate compared to previous studies (Table 3.3) resulted in more time for 8:2 FTOH to leave the sediment-leachate phase in live-spiked microcosms compared to previous studies (Liu et al., 2007; Wang et al., 2005a, 2005b, 2009). Therefore, leachate storage (e.g., in the evaporation pond) and on-site treatment, especially involving aeration (e.g., aerated lagoons, air stripping, etc.), could represent a significant pathway for semi-volatile precursors (e.g., 8:2 FTOH) to enter the atmosphere.  More than 25% of the initially spiked 8:2 FTOH was unaccounted for after 90 d, likely due to unknown metabolites and known metabolites (e.g., 7:3 FTUCA, 7:2 FT ketone and 8:2 FT aldehyde) that could not be quantified due to lack of authentic standards. In addition, irreversible binding of 8:2 FTOH and its metabolites has been previously reported in soil (Wang et al., 2009; Liu and Lee, 2005), which might explain the decreasing mass recovery from the sediment-leachate phase of the live-spiked samples (Figure 3.4a). The overall higher recovery of 8:2 FTOH from sterile samples (Figure 3.4b), compared to live-spiked microcosm (Figure 3.4a) indicates that binding of the organic matter and fluorinated chemicals were likely catalyzed by microbial enzymes (Zhao et al., 2013a; Liu et al., 2010b). Previous studies (Kaestner et al., 2014; Xu and Bhandari, 2003; Dec and Bollag, 1997; Bollag, 1991) demonstrated that oxidoreductase enzyme (e.g., peroxidase, laccase) are able to catalyze covalent bonds between organic pollutants and humic substances. The non-extractable residue of FTCs in soil/sediment has the potential to reduce their toxicity and bioavailability (Kästner and Richnow, 2001). However, with regards to bioremediation techniques, the long-term stability and remobilization potential of the non-extractable residue of FTCs need to be assessed (Kästner and Richnow, 2001).  45   Figure 3.4 Concentrations of the 8:2 FTOH in various sampling media a) live-spiked; b) spiked sterile controls; c) concentrations of 6:2 FTS in spiked live and sterile microcosms. All concentrations are normalized to the initial spiking concentration (~500 nmol/L for 8:2 FTOH and ~3000 nmol/L for 6:2 FTS). Absolute ranges of the duplicate samples (n = 2) are expressed by error bars. Error bars smaller than the symbol height are not visible. The observed average recovery of 90 – 110% in the sterile control microcosm indicates that our extraction method was able to recover 6:2 FTS efficiently from landfill leachate and sediment. Live-spiked microcosms showed a decreasing trend of recovered 6:2 FTS, reaching about 53% after 90 d (Figure 3.4c). This suggests the loss of the parent compound by biotransformation and 46  irreversible adsorption. Considering the formation of biotransformation products, ~35% of the spiked 6:2 FTS was unaccounted for at 90 d, slightly more than for 8:2 FTOH (Figure 3.4a). Overall, the greater biotransformation extent of 6:2 FTS translates into more unaccounted transformation products and their non-extractable bound residue in sediment, compared to 8:2 FTOH.  Biotransformation pathways of 8:2 FTOH and 6:2 FTS Detection of PFCAs and X:3 acids indicates that the biotransformation of 8:2 FTOH and 6:2 FTS follow pathways in the sediment-leachate system similar to those in activated sludge, river sediment and soil; although with different molar yields of the metabolites, as outlined in Tables 3.3 and 3.4. The biotransformation pathways of 8:2 FTOH and 6:2 FTS are summarized based on this study and previous studies (Van Hamme et al., 2013; Wang et al., 2005a, 2005b, 2009, 2011) in Figure 3.5. Biotransformation of 6:2 FTS starts by microbial desulfonation, likely catalyzed by oxygenase enzymes (Van Hamme et al., 2013; Wang et al., 2011), leading to the formation of 6:2 FTOH, which then follows the same alcohol metabolism pathway as for 8:2 FTOH (Figure 3.5). FTOHs are first oxidized to fluorotelomer aldehyde (FTAL) (Wang et al., 2005a) by alcohol dehydrogenase enzyme, which is further oxidized to FTCA, likely by aldehyde dehydrogenase (Cederbaum, 2012). The FTCA is converted to FTUCA through the first defluorination reaction with a dehydrohalogenation mechanism. Dehydrohalogenation involves removal of a halide atom from one carbon atom, with the removal of a hydrogen atom from an adjacent carbon atom, followed by the formation of a double bond (Vogel et al., 1987). This process can be biotic (Wang et al., 2009) or abiotic, as observed for chlorinated insecticides (Gianfreda & Bollag 2002; Nagata et al., 1993; Imai et al., 1989). The significant role of abiotic dehydrohalogenation in the biotransformation of chlorinated solvents at an environmentally relevant pH range is well documented (He et al., 2015; Tobiszewski and Namieśnik, 2012). Formation of FTUCA is a vital branching point in FTC biotransformation, following which the pathway may proceed in two ways, the PFCA pathway or the X:3 acid pathway (Figure 3.5). In the PFCA pathway, FTUCA is biotransformed into FT ketone (Wang et al., 2009) through multiple steps involving defluorination and decarboxylation. The ketone is then converted to secondary FTOHs. PFCAs are formed from 47  secondary alcohols, possibly via multiple enzymatic steps (Kim et al., 2012), which are unknown at this time. The X:3 acid pathway involves defluorination of n:2 FTUCA (n = 6 or 8) to X:3 FTUCA, subsequent reduction of X:3 FTUCA to X:3 FTCA by double bond reductase at the expense of nicotinamide-adenine dinucleotide phosphate (NAD[P]H), which serves as the hydride donor (Huang et al., 2014; Van Hamme et al., 2013). X:3 FTUCA can also enter a “one-carbon removal pathway” leading to the formation of PFCAs, as described by Wang et al., 2012a.  Since biotransformation of X:3 FTCA is known to produce more short-chain PFCAs (Figure 3.5), it might be desirable from the bioremediation point of view to shift the transformation of FTCs in leachate-sediment towards the X:3 pathway. This is especially relevant for long-chain compounds like 8:2 FTOH, whose biotransformation proceeds via the PFCA pathway and primarily produce legacy contaminants like PFOA (Figure 3.3b). To date, the factors controlling the split between the PFCA and X:3 pathways in complex mixed microbial cultures such as landfill leachate are unknown. In addition, lower biotransformation potential of 7:3 FTCA (section 3.3.5) might present a bottleneck in utilizing the X:3 pathway. Further studies are needed to understand the environmental and physiological conditions that enhance bioavailability and biotransformation of 7:3 FTCA (Wang et al., 2012a) during biological leachate treatment.        48  F-(CF2)8-(CH2)2-OH(8:2 FTOH)F-(CF2)6-(CH2)2-SO3H(6:2 FTS)DesulfonationF-(CF2)6-(CH2)2-OH(6:2 FTOH)F-(CF2)n-CH2-CHO(n: 2 FTAL n = 6 or 8)F-(CF2)n-CH2-COOH (n: 2 FTCA)F-(CF2)n-1 -CF=CH-COOH (n: 2 FTUCA)F-(CF2)n-1 -CO-CH3 ([n-1]: 2 FT ketone)F-(CF2)n-1 -CHOH-CH3 ([n-1]: 2 FT secondary alcohol)Loss of one or more -CF2 unitF-(CF2)n-1-CH=CH-COOH([n-1]:3 U acid)F-(CF2)n-1-CH(OH)-CH2-COOH(3-hydroxy-[n-1]:3 U acid)F-(CF2)x -COOH(PFCA; x = 5, 6 for n = 8 and x = 4, 5 for n = 6)F-(CF2)x -COOH(PFCA; x = 4, 5, 6 for n = 6 and x = 6, 7, 8 for n = 8)Loss of one or more -CF2 unitF-(CF2)n-1-CH2-CH2-COOH([n-1]:3 acid) Figure 3.5  Aerobic biotransformation pathways of 8:2 FTOH and 6:2 FTS in sediment-leachate in comparison with previous studies. The metabolites in blue have been reported in previous studies (adapted from Wang et al., 2005a, 2005b, 2009, 2011).   “-HF”     49  3.4 Conclusions and future research Under aerobic conditions, 8:2 FTOH and 6:2 FTS persisted in landfill leachate and sediment (half-life >>30 d). Slower biotransformation led to significant partitioning of 8:2 FTOH to the gas phase, suggesting that landfills may act as secondary sources for semi-volatile FTOHs in the environment. The abundance of PFOA as an 8:2 FTOH biotransformation product suggests that landfills represent a long-term source of legacy contaminants like PFOA in the environment. Transformation of predominantly C4 to C6 PFCAs from 6:2 FTS indicates that the shift towards short-chain fluorotelomer replacements results in a greater abundance of short-chain PFCAs in landfill leachate, followed by their release to the aquatic environment.  The biotransformation of fluorotelomer compounds in leachate was accompanied by the oxidation of organic carbon and conversion of ammonia to nitrate, indicating heterotrophic and autotrophic microbial growth. Knowing the role of various microbial communities could provide a better understanding of the fate of FTCs in various unit processes during biological leachate treatment. It is likely that the dissolved organic carbon of the landfill leachate supplied additional carbon source and retarded the overall 8:2 FTOH and 6:2 FTS biotransformation, which have been characterized as cometabolic in nature. Further studies are needed to elucidate the effect of leachate constituents (e.g., organic carbon, ammonia) on FTCs biotransformation. While soil/sediment-bound residue can potentially reduce toxicity and bioavailability of FTCs, long-term stability and remobilization potential of such residue from contaminated soil (e.g., with subsurface leachate migration) need to be studied. Future research is also necessary to develop effective removal techniques for short-chain PFCAs from landfill leachate to limit their release in the environment.  50  Chapter 4: Effect of Substrate Concentrations on Aerobic Biotransformation of 6:2 Fluorotelomer Sulfonate in Landfill Leachate 4.1 Introduction Fluorotelomer sulfonates (FTSs) are widely used as surfactants and water repellents in various coatings, paints, adhesives, wax and polishes applied as surface treatment agents in numerous consumer products made of paper, wood, leather and textiles (DuPont, 2012). Short-chain (<C7) sulfonates such as 6:2 FTS are increasingly replacing legacy long-chain fluorinated compounds (e.g., perfluorooctane sulfonate (PFOS)) as mist suppressant products, such as Fumetrol 21, in chrome plating (Yang et al., 2014; Poulsen et al., 2011). As 6:2 FTS-containing consumer products are disposed at the landfill at the end of their useful lives, 6:2 FTS have been detected in landfill leachate at a concentration range between 20 and 500 ng/L (Knutsen et al., 2019; Allred et al., 2014; Lang et al., 2017). Previous studies have shown that 6:2 FTS biotransformation can lead to the formation of mixtures of short-chain (C4 – C6) PFCAs and fluorotelomer acids in various environmental media (e.g., river sediment, activated sludge) (Ochoa-Herrera, 2016; Zhang et al., 2016; Wang et al., 2011), including landfill leachate (section 3.3.5 in Chapter 3). Depending on environmental media, the biotransformation rate of 6:2 FTS can vary substantially, with half-lives ranging between less than a week to years (Zhang et al., 2016; Wang et al., 2011) (Table 3.3 Chapter 3). The molar yields of stable products (e.g., PFCAs) also vary based on the biotransformation medium. Short-chain PFCAs are highly mobile (water solubility >20 g/L) (Ateia et al., 2019; Brendel et al., 2018; Vierke et al., 2014; Wang et al., 2011), resulting in their widespread detection in ground and surface water (Banzhaf et al., 2017; Yao et al., 2014; Backe et al., 2013), landfill leachate (Knutsen et al., 2019), ocean (Kwok et al., 2015), with a concentration range of single to several thousands of ng/L (Li et al., 2019). Concerns exist regarding the persistence of short-chain PFCAs (Cousins et al., 2016; Liu and Mejia-Avendaño, 2013), making it challenging to remove them from wastewater and drinking water (Li et al., 2019; Rahman et al., 2014). To limit the release of short-chain PFCAs in the environment, it is important to understand the contribution of secondary sources such as biotransformation of fluorotelomer compounds. 51  Landfill leachate quality and quantity vary greatly depending on waste quantity and type, landfill age, climatic conditions, etc. (Tsarpali et al., 2012; Kjeldsen et al., 2010; Renou et al., 2008). Biotransformation of fluorotelomer compounds are hypothesized to be cometabolic in nature (Lewis et al., 2016; Kim et al., 2012) and linked to growth substrates (e.g., total organic carbon, ammonia). Therefore, variation in the total organic carbon (TOC) and ammonia concentrations in leachate are expected to affect 6:2 FTS biotransformation outcomes (Fischer and Majewsky, 2014; Plósz et al., 2009), which are unknown to date. The overall goal of this study was to investigate the effect of varying concentrations of TOC and ammonia in 6:2 FTS biotransformation and PFCA formation in landfill leachate under aerobic conditions. To achieve this, 6:2 FTS biotransformation experiments were conducted in microcosms containing sediment, from a leachate collection ditch, which was added with deionized (DI) water, diluted landfill leachate or landfill leachate. Spiked 6:2 FTS and biotransformation end products (i.e., PFCAs) were quantified for each experimental condition to understand the effect of landfill leachate constituents on the production of PFCAs. Microbial communities under different experimental conditions were investigated using 16S rRNA sequencing to better understand their response to changing substrate conditions and the potential effect of 6:2 FTS biotransformation. 4.2 Materials and methods  Materials, standards and reagents Landfill sediment and leachate were collected in August 2018, from a leachate collection ditch in a municipal landfill (location kept confidential at the request of the operator), and then used to prepare biotransformation microcosms. The sediment and leachate sampling have been described in section 3.2.2 of Chapter 3. Following the transport of the collected samples to the lab, sediment samples were wet sieved via a 2 mm sieve (ASTM E11 No. 10 size). Sieved sediment was stored overnight with 40 – 50 mm of leachate on top at 4°C and then used for microcosm preparation the next day. The initial physical and chemical characteristics described in Table 4.1 showed overall higher organic and inorganic leachate constituents compared with the leachate used for the 90 days (d) study described in Chapter 3. This is because the preceding 30 d of the sampling day recorded <10 mm rainfall at the landfill site, resulting in a more concentrated in leachate. 52  6:2 FTS and three of its known biotransformation products (perfluorobutanoic acid (PFBA), perfluoropentanoic acid (PFPeA) and perfluorohexanoic acid (PFHxA)) were monitored. Perfluoroheptanoic acid (PFHpA) and perfluorooctanoic acid (PFOA) were also quantified to monitor background biotransformation of longer chain fluorotelomer compounds (e.g., 8:2 fluorotelomer alcohol) already present in the landfill leachate or sediment, which can also contribute towards PFHxA formation (see Section 3.3.5 in Chapter 3). The acronyms, CAS number and suppliers of the analytes of interest are listed in C.1 of Appendix C. HPLC grade ammonium acetate, methanol and acetonitrile were purchased from Fisher Scientific, Canada. Ultrapure water (Type 1) from a Milipore® system was used for all purposes. Oasis weak anion exchange (WAX) (6 cc/30 µm) was obtained from Waters (MA, USA). Millex® syringe filters (0.45 µm, 33 mm diameter) with hydrophilic polyvinylidene fluoride (PVDF) membranes were purchased from MilliporeSigma, Canada. Table 4.1 Initial characterization of sediment and leachate samples (n=3) Sediment Characteristics  pH 7.5 ± 0.2%  (sediment: distilled water 1:2 v/v) Moisture content (%) (wet wt. basis) 80 ± 0.5% Organic content (%) (dry wt. basis) 7 ± 3% Leachate Characteristics  pH 8.3 ± 0.4% Electrical conductivity (mS/m) 28 ± 3% Total organic content (mg/L) 540 ± 4% Alkalinity (mg CaCO3 mg/L) 1870 ± 1% Ammonia (mg/L) 145 ± 2%        Metalsa (mg/L)  Calcium 147 ± 8% Magnesium 55 ± 7% Silicon  30 ± 4%       Anions (mg/L)  Fluoride < 0.5 Chloride 625 ± 1% Sulfate 39 ± 3%       ametals higher than the detection limit (0.5 mg/L) are shown 53   Microcosm preparation Biotransformation microcosms were prepared, using the collected landfill sediment and leachate, in Wheaton glass serum bottles (125 mL) with aluminum crimp-sealed natural rubber stoppers. According to the experimental conditions shown in Table 4.2, each bottle received 15 g of wet sediment, 5 mL mineral media, and 10 mL landfill leachate or deionized (DI) water at various ratios. The preparation of mineral medium has been described in section 3.2.3 in Chapter 3. Sterile control microcosms were autoclaved using two cycles at 121°C for 1 h, with intermittent incubation (20°C for 24 h) between cycles. To prevent microbial growth in the sterile controls during incubation, 0.75 g/L of NaN3 was added to each sterile bottle. NaN3 is commonly used as an inhibitor of bacterial activity.  It is an effective electron transport chain inhibitor for cytochrome oxidase and catalase (Keilin, 1936), that causes chemical asphyxiation of the cell (Winter et al., 2012). The live-spiked and sterile microcosms were dosed with a diluted stock of 6:2 FTS (50 g/L in DI water) to achieve a spiking concentration of ~650 µg/L. The live-controls were dosed with an equal volume of DI water only. All microcosms were placed inside a dark, temperature-controlled (20 °C) orbital shaker at 150 rpm (Innova 4200, New Brunswick Incubator Shaker). Table 4.2 Experimental conditions of 6:2 FTS aerobic biodegradation  Condition name Microcosm Spikea Inhibitor Live-spiked Deionized (DI) water Sediment, DI water, mineral media 6:2 FTS None Diluted leachate Sediment, leachate diluted with DI water (50:50; v/v), mineral media 6:2 FTS None Leachate Sediment, landfill leachate, mineral media 6:2 FTS None Sterile-control Sterile Sterilized sediment, landfill leachate, mineral media 6:2 FTS NaN3 Live-controls DI water control Sediment, DI water, mineral media None None Diluted leachate control Sediment, leachate diluted with DI water (50:50; v/v), mineral media None None Leachate control Sediment, landfill leachate, mineral media None None aSpiking concentration of 6:2 FTS was  ̴ 650 µg/L   Sample collection, preparation and instrumental Analysis For each experimental condition listed in Table 4.2, duplicates microcosms were sacrificially sampled after 0, 1, 4, 7, 14, 30 and 60 days (d). On each sampling day, the headspace concentration 54  of O2 was measured in the sampled live-spiked, live-controls and sterile-controls using a Quantek oxygen analyzer 905 (Grafton, MA, USA), followed by pH measurement of the microcosm slurries. At 30 d, the headspace O2 level in the live-spiked and live-control microcosms approached the recommended minimum level of 10% oxygen saturation (i.e., 2% headspace O2 content) (Mejia-Avendaño et al., 2016; Liu and Mejia-Avendaño, 2013). Therefore, the remaining bottles were reaerated (Figure 4.1a). TOC, ammonia, nitrate and nitrite were also determined on each sampling day, using the dissolved phase (filtered through 0.45 µm filter paper) of the microcosms. To determine 6:2 FTS and PFCAs, 10 mL of microcosm slurry was collected on each sampling day, stored at -17°C and subsequently extracted within two weeks. In addition, a 5 mL sample was withdrawn from each replicate on days 0, 7, 14, 30 and 60, and stored at -80°C for microbial community analysis.  Poly- and perfluoroalkyl substance (PFAS) analysis The PFCAs were extracted from a 5 mL sample added with MPFHxA as surrogate standard, using two cycles of solvent extraction (5 mL 50% (v/v) acetonitrile and methanol). The extracts were diluted with DI water and cleaned up with solid-phase extraction (SPE) using Oasis WAX® following a modified method (USEPA 2011). The SPE extract was dried under a gentle N2 stream, reconstituted with injection solvent (95% aqueous methanol) and spiked with labelled internal standards (M3PFBA and MPFOA) prior to HPLC-MS/MS analysis. Extraction of 6:2 FTS was done using 0.5 mL of sample added with M2_6:2FTS as surrogate standard, and two cycles of solvent extraction. The extract was centrifuged, filtered through a 0.45 um PVDF syringe filter, added with internal standard (M8PFOA) analyzed using HPLC-MS/MS. The detailed extraction and clean-up methods for PFCAs and 6:2 FTS are provided in section C.2 of Appendix C.  The separation of PFASs was done using an Agilent 1200 series HPLC system (Agilent Technologies, CA, USA). 6 µL of sample was injected onto a Waters Xterra MS C18 column (100×2.1 mm, 3.5 µm particle size; Waters Corporation, Milford, MA), preceded by an Xterra MS C18 VanGuard cartridge (30×2.1 mm, 3.5 µm particle size Waters Corporation, Milford, MA). The mobile phase consisted of DI water with 5 mM ammonium acetate, and (B) 95% aqueous methanol with 5 mM ammonium acetate. The detailed solvent gradient and HPLC-MS/MS instrumental conditions are specified in C.3 of Appendix C.  55   Organic and inorganic contents analysis The ammonia, nitrate and nitrite were analyzed by a QuickChem® (Lachat Instrumets, WI, USA) automated ion analyzer. A Phoenix® 8000 TOC analyzer (Teledyne Tekmar, OH, USA) using the UV-persulfate method was used for TOC analysis of the dissolved phase of the microcosms. Fluoride was analyzed by a Dionex™ (ICS-900) ion chromatography (IC) (Dionex Inc., Sunnyvale, CA, USA) system. The separation of anions in IC was done using an IonPac™ AS4A-SC (4 mm × 250 mm) analytical column and IonPac™ AG4A-SC guard column. The mobile phase consisted of 3.5 mM sodium carbonate and 1.0 mM sodium bicarbonate (flow rate of 1.2 mL/min).   DNA extraction, PCR amplification, sequencing and post-sequencing analysis Duplicate microcosm samples on each sampling day (i.e., 0, 7, 14, 30 and 60 d) were mixed together, from which a 5 mL sample was withdrawn for deoxyribonucleic acid (DNA) analysis. DNA extraction and PCR, post-sequencing analyses were performed by Microbiome Insights (Vancouver, Canada). DNA was extracted using MoBio PowerMag Soil DNA Isolation Bead Plate, optimized for the Thermofisher KingFisher™ robot. Bacterial 16S rRNA genes were polymerase chain reaction (PCR)-amplified with dual-barcoded primers targeting the V4 region, according to the protocol of Kozich et al., (2013). Amplicons were sequenced with an Illumina MiSeq using the 300-bp paired-end kit (v.3). Sequences were denoised, taxonomically classified using Greengenes (v. 13_8) as the reference database, and clustered into 97% similarity operational taxonomic units (OTUs) with the mothur software package (v. 1.39.5), following the procedure recommended by Schloss et al., (2009).  Quality control and quality assurance Throughout the sample processing and storage, polypropylene and HDPE were preferred over polytetrafluoroethylene (PTFE) containing materials to eliminate contamination by background fluorotelomer compounds. Duplicate microcosms were used for all analyses including TOC, ammonia, nitrate, nitrite and PFAS. Method blanks were extracted, cleaned up using SPE and analyzed, following the same procedure as the samples to account for any PFAS contamination during sample processing. HPLC-MS/MS runs included quality control (QC) intermediate level calibration standards, solvent blanks injected every 10 – 15 samples to monitor absolute analyte 56  areas, chromatographic retention times and background contamination from the instrument. Duplicate microcosm samples were analyzed at days 0 and 14 for microbial community analysis. The analytes of interest were below detection limits in all procedural and solvent blanks. The accuracy of the QC standards varied in the range of 90 – 110%. The recovery ranges of the labelled surrogates, MPFHxA and M2_6:2 FTS, were 70 – 130% and 85 – 130%, respectively. C4 – C8 PFCAs were quantified in the all live-control microcosms (without 6:2 FTS spike).  4.3 Results and discussion  Microcosm monitoring  The parameters of microcosms monitoring (i.e., headspace oxygen content, pH, TOC, nitrogenous compounds) of 6:2 FTS live-spiked and control microcosms are shown in Figures 4.1 and 4.2. The physical and chemical parameters of each live-spiked microcosm (Figure 4.1) matched well with its corresponding live-control without 6:2 FTS spike (Figure 4.2) and did not show statistically significant differences (α = 0.05). On the other hand, all the monitoring parameters among the three live-spiked conditions showed a statistically significant difference, suggesting that the addition of DI water, diluted leachate and landfill leachate resulted in microcosms with distinct characteristics.  The headspace oxygen depleted in all live microcosms, with the fastest rate being observed in microcosms with landfill leachate (Figure 4.1a), suggesting overall higher microbial activity. In order to ensure aerobic conditions throughout the experimental period, reaeration was provided after 30 d to remaining live microcosms. The initial pH drops of the live microcosms around 7 d coincided with the initial depletion of the TOC (Figure 4.1c) and ammonia (Figure 4.1d). The CO2 produced by aerobic microbial respiration reacts with water to form carbonic acid, which can cause a decrease in pH. Furthermore, neutralization of the hydrogen ions released during ammonia oxidation to nitrate can also reduce the pH.  The initial TOC of the live microcosms varied between 350 to 100 mg/L (Figures 4.1c and 4.2c). The removal of TOC ranged between 50 and 65 % in the DI water and leachate microcosms and 70 – 75% for the diluted leachate microcosms. The depletion of initial ammonia (30 to 110 57  mg/L) (Figures 4.1d) and nitrite contents (Figures 4.1e) were accompanied by an accumulation of nitrate (Figure 4.1f) in live microcosms, suggesting growth of nitrifiers. By 30 d, nearly complete removal of ammonia was observed in live-spiked microcosms (Figure 4.1d). Comparison of the TOC (∽200 mg/L) and ammonia (∽60 mg/L) concentrations revealed that the diluted leachate microcosms had similar substrate concentrations of that in the live microcosms used for 90 d biotransformation study (Figure 3.1) described in Chapter 3.  Figure 4.1 Time variation of a) headspace oxygen content (%), b) pH, c) total organic carbon (TOC; mg/L), d) ammonia (mg N/L), e) nitrite (mg N/L) and f) nitrate (mg N/L) in the dissolved phase of microcosms added with deionized (DI) water, diluted leachate (50:50; leachate: DI water) and leachate. All microcosms were spiked with 6:2 FTS. The absolute differences of duplicate measurements are represented by error bars.  58   Figure 4.2 Time variation of a) headspace oxygen content (%), b) pH, c) total organic carbon (TOC; mg/L), d) ammonia (mg N/L), e) nitrite (mg N/L) and f) nitrate (mg N/L) in the dissolved phase of live-control and sterile microcosms. The live-controls were added with deionized (DI) water, diluted leachate (50:50; leachate: DI water) and leachate, without any 6:2 FTS spike. The sterile microcosms were spiked with 6:2 FTS. The absolute difference of duplicate measurements is represented by the error bars.    59   Microbial community analysis 4.3.2.1 Initial microbial community composition of landfill leachate-sediment at Phylum level The final dataset had quality-filtered reads ranging between of 6×103 and 3.5×104 for various samples. Exhaustive data on the bacterial community were obtained for each sample, as indicated by the sequencing coverage rate that exceeded 99%. The relative abundance of the microbial community at day 0 in live-spiked microcosms at the phylum level is shown in Figures 4.3 (Figure C.1 for live-controls). Proteobacteria (>70%), Bacteroidetes (>10%), Chloroflexi (>1.5%) and Firmicutes (>2%) accounted for >83.5% of the observed phyla at day 0 in all live-spiked microcosms. There was no significant difference (p > 0.05) in the composition of bacterial phyla in the DI water, diluted leachate and leachate microcosms at day 0. The dominance of Proteobacteria, Bacteroidetes and Firmicutes in landfill leachate (Yang and Song, 2019; Stamps et al., 2016; Song et al., 2015; Köchling et al., 2015), landfilled waste (Xu et al., 2017; Wang et al., 2017) and cover soil (Wong et al., 2019; Wang et al., 2017; Mwaikono et al., 2016) have been shown previously. Proteobacteria is the largest and most diverse bacterial phylum that plays a crucial role in nutrient cycling (Newton et al., 2011). All five classes (i.e., Alpha-, Beta-, Gamma-, Delta, Epsilonproteobacteria) (Gupta, 2000) of Proteobacteria were detected, with Beta-, and Epsilonproteobacteria alone constituting 50 – 80% of the relative abundance in day 0 samples. Many genera of Alpha-, Beta- and Gammaproteobacteria are known degraders of aliphatic and aromatic hydrocarbon (Kuppusamy et al., 2016; Sutton et al., 2013; Juhasz and Naidu, 2000; Parales, 2010). Pseudomonas spp., known to biotransform fluorotelomer compound cometabolically, belongs to Gammaproteobacteria (Lewis et al., 2016; Kim et al., 2014; Kim et al., 2012). In addition, known methane-oxidizer from family Methylococcaceae, belonging to Gammaproteobacteria were also present in all microcosms at day 0. The key enzyme used by methane-oxidizers, methane monooxygenase, is capable of oxidizing diverse recalcitrant compounds (e.g., chlorinated solvents) (Knief 2015; Jiang et al., 2010). Firmicutes plays an important role in cellulose degradation to form sugars, whereas, Bacteroidetes metabolize sugars to carboxylic acids, alcohols and carbon dioxide (Yang and Song, 2019; Semaru, 2011).  60   Figure 4.3 Relative abundance of microbial community structures at the phylum level in 6:2 FTS spiked microcosms at day 0. Lower abundance (<1%) taxa are summed and represented as “Others”. Duplicates were analyzed for 0 and 14 d samples. The absolute difference of relative abundances of duplicates samples ranged between 0.02 to 0.1%. 4.3.2.2 Change of microbial community over 60 days While the phylum level composition of the live spiked microcosms changed over time under the three experimental conditions, Proteobacteria dominated the microbial composition and accounted for >63% at 60 d (Figure C.1). Acidobacteria and Actinobacteria increased throughout the experiment, showing 20- and 3-times increase of relative abundance after 60 d (Figure 4.4) compared to day 0, respectively. Acidobacteria is one of the most widespread and abundant soil bacteria phyla (Naether et al., 2012). However, very little is known about their functional role in terrestrial ecosystems due to difficulties associated with the cultivation of Acidobacteria (Kielak et al., 2016). Generally, Acidobacteria is known to thrive under low nutrient conditions, which might explain their higher abundance towards the end of the experimental period. While studies have reported tolerance of Acidobacteria towards several pollutants, for example, petroleum compounds (Abed et al., 2002), p-nitrophenol (Paul et al., 2006), linear alkylbenzene sulfonate (Sanchez-Peinado et al., 2010) and heavy metals (Barns et al., 2007; Ellis et al., 2003), no study has reported pollutant degradation by Acidobacteria (Kielak et al., 2016). Actinobacteria plays an important role in the global carbon cycle by decomposing plant biomass (Lewin et al., 2016). 00.250.50.751DI water Diluted leachate LeachateDay 0Relative abundance OthersBacteria_unclassifiedProteobacteriaFirmicutesChloroflexiBacteroidetes61  Actinobacteria can synthesize a wide array of enzymes including dehydrogenases, peroxidases, monooxygenases and dioxygenases (Donova, 2007), and they have been shown to perform microbial desulfurization (Schmalenberger et al., 2008; Vermeij et. al., 1999). Many genera of Actinobacteria have received special attention as bioremediation candidates due to their biotransformation ability of organic (e.g., alkanes, polycyclic aromatic hydrocarbons (PAHs), phenols, polychlorinated biphenyls (PCBs), halogenated hydrocarbons) (Krivoruchko et al., 2019; Alvarez et al., 2017; Donova, 2007; Schrijver and Mot, 1999). These suggest that Actinobacteria can contribute towards the observed 6:2 FTS biotransformation and PFCA formation in live microcosms.  Figure 4.4 Relative abundance of phylum Actinobacteria and Acidobacteria in a) live-spiked and b) live-control microcosms. Lower abundance (<1%) taxa are summed and represented as “Others”. Duplicates were analyzed for 0 and 14 d samples. The absolute difference of relative abundances of duplicates samples on 0 and 14 ranged between 0 and 0.2%. 62  Ammonia-oxidizer from family Nitrosomonadaceae (genera Nitrosomonas and Nitrosovibrio) and nitrite-oxidizer genera Nitrospira, (family  Nitrospiraceae) were detected in the live microcosms throughout the 60 d. Summation of the known nitrifying bacteria accounted for ~0.4% of relative abundance at day 0, which increased to ~2% around 30 d, followed by a slight decrease to ~1.5% at 60 d in the live-spiked microcosms (Figure 4.5). The increase and subsequent decrease of nitrifying bacteria were primarily due to the growth of ammonia-oxidizers (from family Nitrosomonadaceae) during 30 d and their subsequent reduction due to complete removal of ammonia between 30 and 60 d (Figure 4.1d).   Figure 4.5 Relative abundance of known nitrifying genera in a) live-spiked microcosms and, b) live-control microcosms (without 6:2 FTS spike). 63  4.3.2.3 Alpha diversity of the microbial community The Shannon index increased throughout the experimental period under all live experimental conditions, as shown in Figure 4.6. At 60 d, increases ranging between 30 and 40% were observed for all treatment conditions, compared with day 0. The increase in the Shannon index indicates an increase in richness (i.e., count of species) for a given pattern of evenness (i.e., the closeness of numbers of each species), and/or an increase in evenness for a given richness (Colwell, 2009). Two-way ANOVA (p = 0.05) indicated that sampling time, rather than the treatment conditions, acted as a significant factor in the diversity index change for the live-spiked and live-control microcosms (C.5 and C.6 of Appendix C).   Figure 4.6 Microbial diversity using Shannon index in a) live-spiked microcosms and, b) live-control microcosms (without 6:2 FTS spike). The absolute differences of Shannon index values of duplicates samples at days 0 and 14 ranged between 0.04 to 11%. 64   Biotransformation of 6:2 FTS and PFCA formation 4.3.3.1 Biotransformation of 6:2 FTS in live-spiked microcosms Spiked 6:2 FTS showed a decreasing trend for all three live microcosms (Figure 4.7). The sterile-control microcosms (Figure 4.7) showed no consistent decrease over the experimental period, suggesting that 6:2 FTS was biotransformed in the live-spikes microcosms.  After 60 d, the removal of 6:2 FTS were ~14, 21 and 21% in DI, diluted leachate and leachate microcosm, respectively. The difference in the amount of 6:2 FTS biotransformed under various experimental conditions after 60 d were not statistically significant. As shown in Figure 4.7, little to no change of the spiked 6:2 FTS was observed until 7 d in the live microcosms. The observed delay likely reflected the time required for the relevant microbial communities to grow until the population was large enough to make an observable change in the 6:2 FTS concentration (Knapp and Bromley-Challoner, 2003). For the DI water microcosm, the observable transformation of 6:2 FTS occurred between 7 and 14 d only, and the concentration did not decrease thereafter. However, diluted leachate and leachate microcosms showed an overall decreasing trend of 6:2 FTS until 60 d, suggesting that substrate availability played an important role in 6:2 FTS biotransformation. Microcosms added with leachate contained more ammonia (>60 mg/L; Figure 4.1d), which could have contributed to the observed higher 6:2 FTS biotransformation in the diluted leachate and leachate microcosms (Figure 4.7). The half-lives based on pseudo-first-order kinetics and 7 data points were ~108 d (R2 = 0.53), 90 d (R2 = 0.80) and 92 d (R2 = 0.81), in the DI, diluted leachate and leachate microcosm, respectively.  The half-life of the spiked 6:2 FTS under diluted leachate condition matched closely with that of observed during 90 d biotransformation study (half-life: ∽86 d shown in Table 3.4) described in Chapter 3, likely due to similar substrate concentrations in these microcosms as discussed in section 4.3.1. 65   Figure 4.7 Concentrations of 6:2 FTS in live and sterile microcosms at various sampling days, normalized to initial spiking concentration on day 0. The initial spiking concentration varied between 635 – 670 µg/L (1490 – 1570 nmol/L). The live-spiked microcosms were added with an equal volume of deionized (DI) water, diluted leachate (50:50; leachate: DI water) and landfill leachate. The absolute difference of duplicate measurements is represented with the error bars.  4.3.3.2 Formation of PFCAs in live-spiked and live-control microcosms The concentrations of C4 to C8 PFCAs increased in the live-spiked and live-control (without 6:2 FTS spike) microcosms throughout 60 d and are presented in Figures 4.8 and 4.9, respectively. To make the levels of PFCAs comparable under various treatment conditions, the concentrations of PFCAs were normalized to the mol % of 6:2 FTS, initially spiked in the live-spiked microcosms at day 0. Formation of C4 – C6 PFCAs resulting from 6:2 FTS biotransformation is shown in Figure 4.8a. Similar to the trend observed for 6:2 FTS biotransformation, little to no increase of ∑C4 – C6 PFCAs was observed between days 0 and 7. After 30 d, the ∑C4 – C6 PFCAs increased to ~7, 3.5 and 2 times the initial ∑ C4 – C6 PFCAs for the DI water, diluted leachate and leachate microcosms, respectively (Figure 4.8a). Between 30 and 60 d, the ∑C4 – C6 PFCAs increased by 0204060801001200 1 4 7 14 30 60mol % of initially spiked 6:2 FTSTime (days)DI water Diluted leachateLeachate Sterile66  a further 4, 5.5, 6 times the initial concentrations for the DI water, diluted leachate and leachate microcosms, respectively, showing an overall higher formation of PFCAs in the leachate-added microcosms compared to DI water. It’s likely that the higher growth substrates of leachate-added microcosms (Figures 4.1c to e) supported biomass growth, therefore, having higher relevant degradative enzymes (Tran et al., 2013). The differences among the ∑C4 – C6 PFCAs formed during the three experimental conditions were not statistically significant at 60 d. The PFBA, PFPeA and PFHxA accounted for 11 – 13%, 43 – 50% and 39 – 44% of the final ∑C4 – C6 PFCA content under the three experimental conditions, suggesting that biotransformation of 6:2 FTS acts as a source of short-chain PFCAs in the environment. The spiked microcosms with DI water had the highest ∑C4 – C6 PFCAs followed closely by diluted leachate and leachate by the end of 60 d. A similar trend with higher ∑C4 – C6 PFCAs in DI water microcosm was observed in the live-control microcosms without 6:2 FTS spike as well (Figure 4.9a). Background biotransformation of precursors that were present in the landfill leachate and sediment (section 3.3.3 in Chapter 3) are likely responsible for the observed short-chain PFCA increase. Considering the background concentrations of the short-chain PFCAs in DI water live-control (Figure 4.9a) at 60 d, ~10 mol% of the initially spiked 6:2 FTS, which accounted for ~74% of the biotransformed 6:2 FTS, was converted to C4 – C6 PFCAs in the DI water microcosm. On the other hand, ~51 and 43% of the biotransformed 6:2 FTS were converted to C4 – C6 PFCAs diluted leachate and leachate microcosms at 60 d (Figure 4.9a). Since spiked DI water microcosm showed the lowest 6:2 FTS biotransformation after 60 d, as discussed in section 4.3.3.1, relatively complete biotransformation of 6:2 FTS to PFCAs resulted in this case. In addition, the leachate-added microcosms likely contained other micropollutants commonly found in landfill leachate (e.g., chlorinated aliphatics, higher alkanes, pesticides, phenolic compounds, PCBs, phthalates (Clarke et al., 2015; Oturan et al., 2015; Kjeldsen et al., 2002). Competition of the 6:2 FTS biotransformation intermediates and other micropollutants from leachate for the same active center of an enzyme can cause competitive enzyme inhibition, resulting in reduced biotransformation (Fischer and Majewsky, 2014). This would also explain relatively higher PFCAs formation in spiked diluted leachate microcosm compared to leachate microcosm (Figure 4.8a), despite showing similar 6:2 FTS biotransformation at 60 d (Figure 4.7). These observations suggest that based on the remediation/treatment goal, it might be desirable to maintain low substrate concentrations during biotransformation to facilitate 67  more complete biotransformation of PFCA precursors in landfill leachate. Also, dilution of landfill leachate during the wet season would likely result in a higher formation of PFCAs from fluorotelomer precursors in landfill leachate.  Figure 4.8 Formation of a) C4 – C6 perfluorocarboxylic acids (PFCAs) resulting from biotransformation of 6:2 FTS in microcosms, and b) C7 and C8 PFCAs in live-spiked microcosms added with equal volume of deionized (DI) water, diluted leachate (50:50; leachate: DI water) and landfill leachate; The initial spiking concentration of 6:2 FTS varied between 635 – 670 µg/L (1490 – 1570 nmol/L). The absolute difference of duplicate measurements is represented with the error bars.  Concentrations of C7 and C8 PFCAs (i.e., PFHpA and PFOA) in live-spiked and live-control microcosms are plotted in Figures 4.9a and b, respectively. The observed increases in PFHpA and PFOA likely resulted from biotransformation of precursor compounds (e.g., 8:2 FTOH) already 68  present in the leachate and/or sediment (Benskin et al., 2012). The initial sum of PFHpA and PFOA (ranging between 0.1 to 0.25 mol%) increased to 1.4 to 1.8 mol% (23 to 29 nmol/L) in the live-spiked microcosms at 60 d and were statistically significantly different (95% confidence level) among the treatments. Since landfill leachate already contains PFCAs such as PFHpA and PFOA, microcosms with leachate added showed a higher concentration of PFHpA and PFOA at day 0. By 7 d the microcosm with DI water showed the highest PFHpA and PFOA formation and the trend continued until the end of the experimental period. Similar to the observation for C4 – C6 PFCAs (Figures 4.8a and 4.9a), a higher overall increase of PFHpA and PFOA occurred in DI water microcosm (Figures 4.8b and 4.9b). Therefore, the DI water microcosm provided the most suitable condition for PFCA formation among the three experimental conditions tested within 60 d. The formation of PFOA in leachate-sediment microcosms, from unidentified precursors, suggests that landfills likely act as secondary sources of legacy PFASs in the environment.    Figure 4.9 Formation of a) C4 – C6 PFCAs, and b) C7 and C8 PFCAs in live-control (no 6:2 FTS spike) microcosms added with deionized (DI) water, diluted leachate (50:50; leachate: DI water) and leachate. The absolute differences of duplicate measurements are represented with error bars.  4.4 Conclusions  The effect of substrate (i.e., TOC and ammonia) concentrations on 6:2 FTS biotransformation was studied using aerobic sediment microcosms, added with deionized (DI) water, diluted leachate or landfill leachate. The microbial community analysis using 16S rRNA analysis indicated that phylum Proteobacteria dominated the bacterial composition, while Acidobacteria and Actinobacteria increased in landfill leachate-sediment microcosms throughout 60 d. Many genera from Proteobacteria and Actinobacteria are capable of synthesizing wide ranges of enzymes and likely play an important role in 6:2 FTS biotransformation. Overall, higher biotransformation of 6:2 FTS was observed in microcosms with added leachate, compared to DI water microcosms, likely reflecting the substrate dependency of 6:2 FTS biotransformation. However, substrate 70  limiting conditions in DI water microcosm resulted in greater formation of short-chain (C4 – C6) PFCAs from 6:2 FTS and C7 – C8 PFCAs from unidentified precursors, compared with leachate-added microcosms. This suggests that dilution of landfill leachate, for example through precipitation, likely results in reduced 6:2 FTS biotransformation and increased PFCAs formation, compared to dry seasons. The landfill leachate used in this study can be considered diluted (TOC <1000 mg/L) compared with the leachate from landfill located in the arid climate. Therefore, further investigation into the effect of highly concentrated leachate on 6:2 FTS biotransformation is recommended. The key enzyme (i.e., ammonium monooxygenase) should be measured to confirm the positive effect of ammonia-oxidizers on 6:2 FTS biotransformation. Future studies are needed to elucidate the relationship between heterotrophic and autotrophic biomass growth and 6:2 FTS biotransformation, to develop a cometabolic model that can predict the removal of parent compounds and PFCAs generation in leachate biological treatment systems.      71  Chapter 5: Role of Microbial Communities in the Formation of Perfluorocarboxylic Acids from 6:2 Fluorotelomer Sulfonate in Leachate 5.1 Introduction Fluorotelomer sulfonates (FTSs) belong to a class of polyfluorinated compounds, widely used in consumer products as surface treatment agents (DuPont, 2012). FTSs are manufactured by telomerization of perfluoroalkyl iodides (Field and Seow, 2017; Krafft and Riess, 2015; Buck et al., 2011). The fluorotelomer chain consists of an even-numbered fluorinated (n) and two non-fluorinated carbons, designated as n:2. One of the most common FTS compounds is the 6:2 compound (n-C6F13CH2CH2SO3-), marketed under the name Capstone™ FS-17 (Yang et al., 2014). Capstone™ is used as a repellent and surfactant in paints, coatings, adhesives, waxes and polishes, applied to various substrates, including paper, wood, metal, textiles and leather (DuPont, 2012). Historically, Zonyl FS-62 and Zonyl TBS polymer coating also contained 6:2 FTS (Field and Seow, 2017; Yang et al., 2014) used in inkjet printing, but they were discontinued in 2014. In recent years, the use of 6:2 FTS as a short-chain alternative to long-chain (>C7) fluorinated compounds (e.g., perfluorooctane sulfonate (PFOS)) is increasing. For example, mist suppressant products used in chrome plating, such as Fumetrol 21, contain 6:2 FTS as a PFOS replacement (Yang et al., 2014; Poulsen et al., 2011).  Many 6:2 FTS-containing consumer products end up at landfills following their use. Recent studies (Lang et al., 2017; Allred et al., 2014; Huset et al., 2011) have shown that leachate (percolating rainwater through waste) from municipal landfills contains many poly- and perfluoroalkyl substances (PFASs), including 6:2 FTS (20 to 500 ng/L). Biotransformation of 6:2 FTS can lead to formation of short-chain (C4 – C6) perfluorocarboxylic acids (PFCAs) under aerobic condition (Field and Seow, 2017), which are subsequently released to the environment (Pan et al., 2016; Arvaniti et al., 2015) due to the lack of an effective removal technique. 6:2 FTS and short-chain PFCAs are frequently detected in drinking water, influent and effluent of municipal wastewater treatment plants (WWTP), freshwater, marine and surface waters, and urban runoff at concentrations below the limit of quantification to several thousand ng/L (Ateia et al., 2019; Field and Seow, 2017). In addition, subsurface migration of leachate from unlined landfills 72  could release PFCAs to the groundwater (Hepburn et al., 2019). Short-chain PFCAs are extremely persistent (Cousins et al., 2016; Liu and Avendaño, 2013) and highly mobile (predicted log Kow of the neutral form: 2.82–4.6, water solubility > 20 g/L, log Koc: 2.7–3.6) (Ateia et al., 2019; Brendel et al., 2018; Vierke et al., 2014; Wang et al., 2011), making their removal during drinking water treatment challenging (Li et al., 2019; Rahman et al., 2014). In addition, their long-term exposure effects to human, wildlife and biota are unknown (Brendel et al., 2018).  Biotransformation of 6:2 FTS has been studied in various mixed microbial cultures (e.g., river sediment, activated sludge) (Ochoa-Herrera, 2016; Zhang et al., 2016; Wang et al., 2011) and pure microbial cultures (Shaw et al., 2019; Van Hamme et al., 2013; Key et al., 1998). The biotransformation starts with desulfonation of 6:2 FTS to form 6:2 fluorotelomer aldehyde (6:2 FTAL). 6:2 FTAL is further oxidized to 6:2 fluorotelomer carboxylic acid (6:2 FTCA), which is converted to 6:2 unsaturated fluorotelomer carboxylic acid (6:2 FTUCA) through a common pathway (Figure 3.5 in Chapter 3). Depending on the microbial strain, the presence of reducing energy and substrates (Lewis et al., 2016; Kim et al., 2014; Kim et al., 2012), 6:2 FTUCA can follow two major lower pathways leading to the formation of 5:3 FTCA and perfluorocarboxylic acids (PFCAs) (e.g., perfluorobutanoic acid (PFBA), perfluoropentanoic acid (PFPeA), perfluorohexanoic acid (PFHxA)) (Zhang et al., 2016; Shaw et al., 2019; Van Hamme et al., 2013; Wang et al., 2011). Wide variations in 6:2 FTS transformation rate (t1/2: <7 days to more than years) and the amount of PFCA formed have been reported under aerobic conditions in different environmental media (Zhang et al., 2016; Wang et al., 2011). This suggests that the extent of 6:2 FTS biotransformation and PFCA production are largely determined by the various microbial communities present in the engineered (e.g., biological treatment processes) and natural systems.  High ammonia (hundreds to several thousand mg/L), high alkalinity (hundreds to several thousand mg CaCO3/L) and organic carbon (tens to several thousand mg/L) of landfill leachate (Gao et al., 2015; Renou et al., 2008) promote oxidation of organic carbon and nitrification in landfill leachate under aerobic conditions. During nitrification ammonium is first oxidized to nitrite by ammonia-oxidizing bacteria (AOB); nitrite is then oxidized to nitrate by nitrite-oxidizing bacteria (NOB) (Tran et al., 2013) (Figure 5.1). While both heterotrophic and autotrophic (i.e., AOB and NOB) can express non-specific enzymes that can break down micropollutants (e.g., 73  pharmaceuticals and personal care products, pesticides, chlorinated solvents etc.) (Fischer and Majewsky, 2014; Tran et al., 2013; Khunjar et al., 2011), increased micropollutant biotransformation by AOB producing ammonia monooxygenases (AMO) have been reported (Fernandez-Fontaina et al., 2016; Helbling et al., 2012; Keener and Arp, 1993) (Figure 5.1). Monooxygenase enzymes incorporate one oxygen atom into the substrate. Due to their functional diversity, monooxygenases can catalyze desulfurization, dehalogenation, denitrification, ammonification and hydroxylation (Karigar and Rao, 2011). Higher AMO activity has also been linked to higher fluorotelomer alcohol (FTOH) degradation and PFCA production in activated sludge (Yu et al., 2018a; 2018b). However, very little is known regarding the role of microbial communities (e.g., heterotrophic, autotrophic) in 6:2 FTS biotransformation and PFCA production in landfill leachate.  Figure 5.1 Conceptual model of nitrification, showing ammonia and nitrite oxidation  One of the most common approaches for understanding the contributions of various microbial communities to pollutant degradation is through inhibition studies. Allylthiourea (ATU) depletes copper ions from the active center of AMO through chelation and acts as a specific inhibitor to AMO (Men et al., 2017). ATU has been widely used to investigate the role of nitrifying bacteria in biotransformation of micropollutants in complex microbial communities (e.g., activated sludge, soil) (Fernandez-Fontaina et al., 2016; Men et al., 2017; Sathyamoorthy et al., 2013; Khunjar et al., 2011; Roh et al., 2009; Tran et al., 2009).  74  The overall goal of this study is to determine the role of heterotrophic (HET) and nitrifying bacteria (i.e., AOB, NOB) in 6:2 FTS biotransformation and PFCA formation in landfill leachate. To achieve this, 6:2 FTS biotransformation experiments were conducted with selected substrates (glucose, ammonia and nitrite) and ammonia-inhibitor (ATU) to stimulate the growth of heterotrophic and nitrifying organisms. 6:2 FTS, stable transformation products (i.e., PFCAs) and fluoride were quantified under each experimental condition to understand the biotransformation of the parent compound, production of PFCAs and defluorination under specific substrate conditions. Differences in microbial communities under different experimental conditions were investigated using 16S rRNA sequencing to better understand the effect of 6:2 FTS and PFCA on microbial population and vice versa. 5.2 Materials and method  Sediment collection and inoculum preparation The sediment used for inoculum preparation was collected from a leachate collection ditch in a municipal landfill whose location is kept confidential as requested by the operator. The sediment and leachate were collected as specified in section 3.2.2 of Chapter 3. Sediment and leachate samples were transported to the lab within 1 hour of collection. Initial physical and chemical properties of the sediment and landfill leachate are provided in Table 4.1 in Chapter 4. Collected sediment was wet sieved via a 2 mm sieve opening (ASTM E11 No. 10 size) and aerated for 2 days with ambient air in the laboratory to remove dissolved organic compounds. The sediment was then allowed to settle, and the supernatant was decanted. The settled sediment was washed twice using pH 7.2 phosphate buffer to remove dissolved organic and nitrogenous compounds (i.e., NH4+, NO2-, and NO3-) (Yu et al., 2018). After washing, the sediment was collected by centrifuging at 1000 rpm for 5 min and the washed sediment was resuspended in mineral media (OECD 1992). The resuspended sediment was kept in aerated condition for a day using ambient air at room temperature (23±2°C) to allow for depletion of background TOC, NH4+ and NO2-.  Standards and reagents The monitored fluorinated compounds included PFBA, PFPeA, PFHxA, PFHpA, PFOA and 6:2 FTS. Their acronyms, CAS number and suppliers are listed in section C.1 of Appendix C. HPLC 75  grade methanol, acetonitrile, ammonium acetate and acetic acid were purchased from Fisher Scientific, Canada. For all purposes, ultrapure deionized water (DI) from a Milipore® system was used. Oasis weak anion exchange (WAX) (6 cc/30 µm) cartridges were obtained from Waters (MA, USA). N-Allylthiourea (ATU) (98%) was purchased from Sigma-Aldrich (Ontario, Canada). Sodium bicarbonate (99%), sodium azide (99%), ammonium sulfate (99%), glucose and sodium nitrite (>95%) were purchased from Thermo Fisher Scientific (MA, USA).  Microcosm preparation The inocula used for all microcosm preparation contained TOC, NH4+ and NO2- at concentrations of ~60 mg/L, <1 mg/L and <1 mg/L, respectively. The inocula were then spiked with 6:2 FTS stock (5 mg/mL in DI) to achieve a concentration of 750 µg/L (~1756 nmol/L) and were allowed to mix using a magnetic stirrer. The spiked inocula were aliquoted into 300 mL Wheaton™ glass bottle with glass stopper to which was added an appropriate substrate, alkalinity, inhibitor and brought to 35 mL by adding DI as per experimental conditions (see Table 5.1). All sample bottles were continuously agitated at 150 rpm on a temperature-controlled orbital shaker at 20 (±1)°C in the dark (Innova 4200, New Brunswick Incubator Shaker). Total suspended solids and volatile suspended solids of the inocula were 1.5 ±0.1 g/L and 0.6 ±0.01 g/L, respectively.  Four conditions with estimated microbial activities heterotrophic (HET), and heterotrophic and nitrifier (HET+AOB+NOB), nitrifier (AOB+NOB) and nitrite-oxidizing bacteria (NOB) were tested using ATU as AOB inhibitor and appropriate substrate (Table 5.1). Sodium bicarbonate was added as alkalinity (~800 mg CaCO3/L) and a source of inorganic carbon for nitrifiers in all microcosms. Ammonia (100 mg/L) was also added to all microcosms to check for the effectiveness of ATU as an AOB inhibitor. The substrate concentrations are based on TOC and NH4+-N concentration of landfill leachate as described in Chapters 3. Sterile control microcosms were prepared by two cycles of autoclaving (at 121°C for 1 h) with intermittent incubation at temperature-controlled shaker (20°C for 24 h) between cycles. NaN3 (0.75 g/L) was added to each sterile bottle to prevent microbial growth. Live microcosms of identical HET, HET+AOB+NOB, AOB+NOB and NOB experimental runs were conducted without 6:2 FTS spike (live-controls) to monitor background PFAS concentration change.  76   Sample collection, preparation and instrumental analysis On each sampling day (days 0, 1, 2, 4, 7 and 10), the pH and dissolved oxygen (DO) level were measured in two bottles for each treatment condition using a Hach HQ30D Portable DO meter. For microbial analysis, 5 mL samples were withdrawn from each microcosm and stored at -80°C. For PFAS analysis, a 10 mL sample was collected and stored at -17°C until further sample preparation. Another 10 mL sample was withdrawn, centrifuged (5000 rpm for 10 min) and filtered through a 0.45 µm cellulose membrane filter (Fisherbrand™). The collected filtrate was stored at -14°C and used for TOC, NH4+, NOx and F- analyses within one week.  Table 5.1 Conditions of the microcosms containing alkalinity (800 mg/L CaCO3), ammonia (100 mg/L) Condition name Spike Inhibitor Glucose (mg C/L) NO2- (mg N/L) HETa 6:2 FTS ATU 200 0 HET+AOB+NOBa 6:2 FTS None 200 0 AOB+NOB 6:2 FTS None 0 0 NOB 6:2 FTS ATU 0 100 Sterile control 6:2 FTS Sodium azide 0 0 HET live control None ATU 200 0 HET+AOB+NOB live control None None 200 0 AOB+NOB live control None None 0 0 NOB live control None ATU 0 100 aHET: heterotrophic; AOB: ammonia-oxidizing bacteria; NOB: nitrite-oxidizing bacteria; ATU: allylthiourea   5.2.4.1 PFAS analysis For PFCAs extraction, a 5 mL sample was added with MPFHxA as surrogate standard and solvent-extracted with acetonitrile and methanol (50:50; (v/v)). The diluted extract was cleaned up by solid-phase extraction (SPE) using Oasis WAX® following a modified method (USEPA 2011). The cleaned-up extract was dried under a gentle nitrogen stream, reconstituted with injection solvent (95% aqueous methanol), spiked with internal standard (M3PFBA and M8PFOA) and 77  analyzed using LC-MS/MS. 6:2 FTS was extracted from 0.5 mL of sample added with M2_6:2FTS as surrogate standard, using two cycles of solvent extraction with methanol. Filtered liquid extracts were aliquoted in LC vials, added with internal standard (M8PFOA) and analyzed using LC-MS/MS. The detailed extraction and clean-up methods for PFCAs and 6:2 FTS are provided in section C2 of Appendix C of SI.  An Agilent 1200 series HPLC system (Agilent Technologies, CA, USA) separated the 6:2 FTS and PFCAs. 6 µL of sample was injected onto a Waters Xterra MS C18 column (100×2.1 mm, 3.5 µm particle size; Waters Corporation, Milford, MA), preceded by an Xterra MS C18 guard column (30×2.1 mm, 3.5 µm particle size) from the same manufacturer. Both columns were maintained at 50°C. The mobile phase consisted of (A) water with 5 mM ammonium acetate, and (B) 95% aqueous methanol with 5 mM ammonium acetate. Solvent gradient and detailed instrumental conditions are specified in section C.3 of Appendix C.  5.2.4.2 Organic and inorganic content analyses Fluoride (F-) was analyzed using Dionex™ (ICS-900) ion chromatography (IC) (Dionex Inc., Sunnyvale, CA, USA) system equipped with an IonPac™ AS4A-SC (4 mm × 250 mm) analytical column and IonPac™ AG4A-SC (10-32) guard column. The mobile phase consisted of 3.5 mM sodium carbonate and 1.0 mM sodium bicarbonate at a flow rate of 1.2 mL/min. The instrumental detection limit of fluoride was 0.025 mg/L. Total organic carbon (TOC) was analyzed using a Phoenix® 8000 TOC analyzer (Teledyne Tekmar, OH, USA) using the UV-persulfate method. Ammonia, nitrate and nitrite (NOx) were analyzed using an automated ion analyzer (QuickChem® 8000, Lachat Instruments, WI, USA).  Quality control and quality assurance To avoid background contamination from fluorotelomer compounds, the use of polytetrafluoroethylene (PTFE) containing material during any stage of sample preparation was eliminated. Instead, HDPE and polypropylene (PP) were used. During the HPLC-MS/MS analysis, quality control intermediate level calibration standards containing native and internal standards were injected every 10 – 15 samples to monitor the absolute analyte areas, the ratio of native analyte to the internal standard area, and chromatographic retention times. Solvent blanks were 78  injected every 10 – 15 samples to monitor any background contamination from the instrument. Procedural blanks were included to account for any contamination during sample preparation. The instrumental detection limit of the analytes varied between 0.5 and 1.0 ng/mL (C.4 of Appendix C). A coefficient of determination (R2) of >0.99 was deemed acceptable for the calibration of the PFASs. Accuracy of the quality control calibration standards varied between 90 and 110%. The recovery ranges of labelled surrogate M2_6:2 FTS and MPFHxA were 90 – 125% and 72 – 120%, respectively. All analytes of interest were below their detection limits in the solvent and procedural blanks. In addition, PFOA and PFHpA were quantified in the all live-control microcosms (without 6:2 FTS spike) to account for background precursor biotransformation.  PCR amplification, sequencing and post-sequencing analysis High-fidelity Phusion® polymerase was used for the amplification of marker genes. To check for carry-over inhibition or a high concentration of DNA, 1:1 and 1:10 dilutions were tested, and the polymerase chain reaction (PCR) products were run on gels for verification. PCR was done with dual-barcoded primers (Kozich et al., 2014) targeting either 16S V4 regions for bacteria using the Illumina MiSeq platform. The barcoding strategy enabled multiplexing up to 384 samples per run. PCR products were verified visually by running a representative subset of samples on a gel. Samples with failed PCRs (or spurious bands) were re-amplified by optimizing PCR conditions. The PCR reactions were cleaned-up and normalized using the high-throughput SequalPrep 96-well Plate Kit. Samples were then pooled to make one library, quantified accurately with the KAPA qPCR Library Quant kit. Intermediate analysis files from MiSeq in FASTQ format were quality-filtered and clustered into 97% similarity operational taxonomic units (OTUs) using the Mothur software package (Schloss et al., 2009). High-quality reads were classified using the Greengenes reference database. A consensus taxonomy for each OTU was obtained. OTU abundances were aggregated into taxonomies.   Data analysis The OTU table was imported into R for statistical analyses using Vegan package. Alpha diversity, the mean species diversity in a community (Whittaker, 1972), was calculated using the Shannon and Simpson index. These indices are a mathematical measure of species diversity accounting for both abundance and evenness of the species present. Species richness (number of different species 79  represented in a community) was estimated using richness indices (Abundance-Richness Coverage Estimator (ACE and Chao1). The Hutcheson t-test (Gardener, 2012) was used to compare Shannon diversity of two community samples. The microbial diversity and relative abundance datasets from the various samples were visualized by plotting via Microsoft Excel 2016. All statistical tests were conducted at a 95% confidence level (significance level = 0.05). Any OTU unit was considered putative contaminants and was removed if their mean abundance in controls reached or exceeded 25 % of their mean abundance in samples. 5.3 Results and discussion  Microcosm monitoring and defluorination  The physical and chemical parameters (i.e., DO, pH, TOC, nitrogenous compounds) of live-spiked and control (live-control and sterile) microcosms are presented in Figures 5.1 to 5.5. Overall, the physical and chemical parameters of live-controls without the 6:2 FTS spike agreed well with the corresponding spiked conditions and did not show statistically significant differences. 5.3.1.1 Dissolved oxygen (DO) The addition of glucose and nitrite, readily biodegradable substrates, resulted in immediate depletion of DO in live HET and NOB condition, as shown in Figures 5.2a to 5.2d, respectively. Subsequent to the decrease, the DO level in HET and NOB increased following the depletion of the substrate (Figures 5.2a and 5.2b), likely through the exchange from the headspace air. However, DO level in HET+AOB+NOB and AOB+NOB showed a decreasing trend throughout the experimental period likely due to the activity of nitrifiers (Figure 5.2a). About 3.3 mg of O2 was consumed per mg of NH4+ removed, which would explain the decreasing DO level in the presence of ammonia oxidation. The DO level in sterile controls did not show any noticeable trend throughout the experimental period (Figure 5.2c) suggesting the absence of microbial activities. DO concentration was >5 mg/L in all experimental conditions at all sampling points, indicating aerobic conditions in the microcosms. 80   Figure 5.2 Time variation of dissolved oxygen (DO; mg/L) in a) live-spiked microcosms with heterotrophic only (HET) and heterotrophic and autotrophic (HET+AOB+NOB); b) live-spiked microcosms with nitrite-oxidizing bacteria (NOB) only and ammonia- and nitrite-oxidizing (AOB+NOB); c) sterile and live-control (no 6:2 FTS spike) microcosms with HET and HET+AOB+NOB growth; d) live-control microcosms with NOB and AOB+NOB growth. The absolute difference of duplicate measurements is represented with the error bars.   5.3.1.2 pH There was no noticeable change in pH between day 0 and day 10 for HET and NOB conditions as shown in Figures 5.3a to 5.3d. On the other hand, the initial pH of 8 dropped to ~7 after 10 d in the presence of ammonia oxidation in HET+AOB+NOB and AOB+NOB conditions (Figures 5.3a and 5.3b). About 6.708 mg HCO3-/mg NH4+ is consumed to neutralize the hydrogen ions released during ammonia oxidation (Grady et al., 2011), resulting in the observed pH drop.   81   Figure 5.3 Time variation of pH in a) live-spiked microcosms with heterotrophic only (HET) and heterotrophic and autotrophic (HET+AOB+NOB); b) live-spiked microcosms with nitrite-oxidizing bacteria (NOB) only and ammonia- and nitrite-oxidizing (AOB+NOB); c) sterile and live-control (no 6:2 FTS spike) microcosms with HET and HET+AOB+NOB growth; d) live-control microcosms with NOB and AOB+NOB growth. The absolute difference of duplicate measurements is represented with the error bars.   5.3.1.3 Substrates  The depletion of glucose added to HET and HET+AOB+NOB conditions (Figures 5.4a and 5.4c) suggests heterotrophic growth. The AOB+NOB and NOB microcosms also had ~70 mg/L of TOC at the beginning of the experiment from the background organic content of the inoculum (Figures 5.4b and 5.4d). While the TOC did not show much variation in AOB+NOB microcosms throughout the experiment, NOB microcosms showed a slight decrease towards the end of 10 d (Figures 5.4b and 5.4d), indicating possible heterotrophic microbial growth. 82   Figure 5.4 Time variation of total organic carbon (TOC; mg/L) in the dissolved phase of a) live-spiked microcosms with heterotrophic only (HET) and heterotrophic and autotrophic (HET+AOB+NOB); b) live-spiked microcosms with nitrite-oxidizing bacteria (NOB) only and ammonia- and nitrite-oxidizing (AOB+NOB); c) sterile and live-control (no 6:2 FTS spike) microcosms with HET and HET+AOB+NOB growth; d) live-control microcosms with NOB and AOB+NOB growth. The absolute difference of duplicate measurements is represented with the error bars.   Added ammonia of HET+AOB+NOB and AOB+NOB was depleted (Figures 5.5a to 5.5d) with a concurrent increase of NO3- (Figures 5.6a and 5.6b) suggesting the growth of AOB and NOB under these conditions. In contrast, ammonia concentration did not decrease in HET and NOB microcosms (Figures 5.5a and 5.5b), where ATU was added, indicating the effectiveness of ATU as an AOB inhibitor. The ammonia content of sterile controls microcosms (Figure 5.5c) did not decrease throughout the experimental period, indicating a lack of microbial growth. As opposed to TOC (i.e., glucose), whose depletion was noticeable from day 1 (Figure 5.4a), ammonia depletion became apparent after 2 days (Figure 5.5a). This is due to a relatively lower yield of autotrophs (~ 0.166 mg VSS/ mg of NH4+ removed) compared to heterotrophs (~ 0.5 mg VSS/mg of glucose removed) (Grady et al., 2011).  83   Figure 5.5 Time variation of ammonia (NH4+; mg N/L) in the dissolved phase of a) live-spiked microcosms with heterotrophic only (HET) and heterotrophic and autotrophic (HET+AOB+NOB); b) live-spiked microcosms with nitrite-oxidizing bacteria (NOB) only and ammonia- and nitrite-oxidizing (AOB+NOB); c) sterile and live-control (no 6:2 FTS spike) microcosms with HET and HET+AOB+NOB growth; d) live-control microcosms with NOB and AOB+NOB growth. The absolute difference of duplicate measurements is represented with the error bars.    84   Figure 5.6 Time variation of nitrite and nitrate (mg N/L) in the dissolved phase of a) live-spiked microcosms with heterotrophic only (HET) and heterotrophic and autotrophic (HET+AOB+NOB); b) live-spiked microcosms with nitrite-oxidizing bacteria (NOB) only and ammonia- and nitrite-oxidizing (AOB+NOB); c) sterile and live-control (no 6:2 FTS spike) microcosms with HET and HET+AOB+NOB growth; d) live-control microcosms with NOB and AOB+NOB growth. The absolute difference of duplicate measurements is represented with the error bars.   5.3.1.4 Fluoride concentration The fluoride concentrations of all microcosms after 10 days were below the instrumental detection limit (0.025 mg/L). Considering the initial concentration of 6:2 FTS to be ~750 µg/L, it can be concluded that little or no defluorination (<5% of the total fluoride content) occurred under any experimental condition, indicating persistence of the end biotransformation products (i.e., PFCAs). 85   Microbial Community Analysis 5.3.2.1 Initial microbial community composition  The sequencing coverage rate exceeded 99% (microbial population) for each sample, indicating that exhaustive data on the microbial community were obtained by MiSeq. The sample at day 0 yielded 156 classified phyla in the domain Bacteria. Figure 5.7 shows that Proteobacteria (>59%) and Bacteroidetes (>28%) were the dominant phyla on day 0. Chlorobi, Chloroflexi and Firmicutes constituted altogether ~7% of the observed phyla. These phyla have been observed in landfill soil (Wang et al., 2017), landfill refuse (Xu et al., 2017) leachate and sludge (Song et al., 2015). Proteobacteria and Bacteroidetes are also frequently found to be the dominant bacterial phyla in marine ecosystems (Steven et al., 2005) and are thought to play important roles in organic matter degradation and the carbon cycle (Newton et al., 2011). Among the five classes of Proteobacteria (Gupta, 2000), three (i.e., Alpha-, Beta- and Gammaproteobacteria) were observed in this study. Previous studies (Sutton et al., 2013; Greer et al., 2010; Parales, 2010; Van Beilen and Funhoff, 2007) have shown that many genera of Alpha-, Beta- and Gammaproteobacteria can utilize various aliphatic and aromatic compounds. Strains of Pseudomonas spp., which belongs to Gammaproteobacteria, were able to degrade fluorotelomer compounds metabolically in the presence of substrates (Lewis et al., 2016; Kim et al., 2014; Kim et al., 2012). Pseudomonas spp. was also present in this study at day 0, suggesting that it might play a role in the biotransformation of 6:2 FTS. Bacteroidetes and Firmicutes specialize in the hydrolysis of polymeric organic matter (e.g., cellulose, starch, protein) (Fernández-Gomez et al., 2013; Cottrell and Kirchman, 2000; Li et al., 2009), the first step in anaerobic degradation in landfills (Semrau, 2011).  86   Figure 5.7 Relative abundance of microbial community structures at the phylum level in the live-spiked and live-control microcosms. Lower abundance (<0.5%) taxa are summed and represented as “Others”. (HET: heterotrophic, AOB: ammonia-oxidizing bacteria, NOB: nitrite-oxidizing bacteria)  5.3.2.2 Microbial community composition at day 7 Proteobacteria and Bacteroidetes dominated the microbial composition for various experimental runs, even at 7 d (Figure 5.7). Pseudomonas from Proteobacteria was present in all samples at day 7 at a relative abundance <0.03%. Slightly higher abundance of Pseudomonas was present in HET+AOB+NOB microcosms, (~0.1%). Actinobacteria dominated the HET+AOB+NOB condition at 7 d (10 – 18 %), as opposed to <1% relative abundance under the rest of the experimental conditions (data shown in section D1 of Appendix D). More than 10 known genera of Actinobacteria were found in HET+AOB+NOB condition, the most abundant being Arthrobacter (>17%) (Figure 5.8). Other genera from Actinobacteria phylum included Acidimicrobiales, Leucobacter, Pimelobacter and Mycobacterium, altogether representing <1% 87  of relative abundance. Among these, Mycobacterium (<0.01%) is a known 6:2 FTOH degrader (Kim et al., 2014). A vermicompost bacterial isolate of genus Gordonia (Shaw et al., 2019; Van Hamme et al., 2013) known to metabolize 6:2 FTS as a sole sulfur source in sulfur-limiting condition, was not detected in any of the samples. Considering the leachate sulfate concentration of ~40 mg/L (Table 4.1 in Chapter 4), it is unlikely that 6:2 FTS would be metabolized as sulfur source in the landfill environment. Ecological factors, such as nutrients, pH, dissolved oxygen and organic matter content can greatly influence the abundance and genera of Actinobacteria (Jiang et al., 2016); this could explain the observed higher abundance of Actinobacteria during the HET+AOB+NOB experimental run. Actinobacteria can synthesize a wide range of enzymes including dehydrogenases, peroxidases, monooxygenases and dioxygenases (Donova, 2007). Consequently, degradation of synthetic compounds such as organochlorine pesticides, linear and branched alkanes C2 – C30, polycyclic aromatic hydrocarbons (PAHs), phenols, polychlorinated biphenyls (PCBs), aromatic acids, halogenated hydrocarbons (Krivoruchko et al., 2019; Donova, 2007; Schrijver and Mot, 1999) by many genera from Actinobacteria are well documented.  Figure 5.8 Relative abundance of phylum Actinobacteria at the genus level in live-spiked and live-control microcosms. Phylum level relative abundance of Actinobacteria varied between 0.3 to 0.8% for all samples, except HET+AOB+NOB, which showed 10 – 18% Actinobacteria at day 7. (HET: heterotrophic, AOB: ammonia-oxidizing bacteria, NOB: nitrite-oxidizing bacteria).  88  5.3.2.3 Nitrifying bacteria Known nitrifying genera Nitrosomonas, Nitrosovibrio belonging to Phylum Proteobacteria were identified in the day 0 sample (Figure 5.9). Among the known NOBs in the environment, only Nitrospira, from phylum Nitrospirae, was detected at the beginning of the experiment. The taxonomic classification of the known nitrifying genera in all samples are provided in section D2 of Appendix D. The summation of the known nitrifying bacteria increased more than 3-fold at 7 d compared to day 0, under the HET+AOB+NOB and AOB+NOB conditions, mainly due to growth of ammonia-oxidizers from the Nitrosomonadaceae family. Nitrospira also slightly increased (>38%) under NOB condition. The abundance of Nitrospira over other known NOBs (such as Nitrobacter) likely occurred because, Nitrospira are adapted to live under a substrate limiting condition (Nogueira and Melo, 2006). Altogether, the known nitrifying bacteria contribute <2 % of the relative abundance at the end of 7 days in live microcosms. In addition to the bacteria detected here, there could be other unknown bacteria contributing towards nitrification. For example, mixotroph organisms, which can use heterotrophic and autotrophic modes of nutrition simultaneously (Crane and Grover, 2010) are able to oxidize ammonia (Kouki et al., 2011).     89   Figure 5.9 Relative abundance (%) of known nitrifying bacteria at the genus level identified through 16s RNA sequencing in live-spiked microcosms. (HET: heterotrophic, AOB: ammonia-oxidizing bacteria, NOB: nitrite-oxidizing bacteria)  5.3.2.4 Alpha diversity and richness of the microbial community The diversity and richness indices (see D.3 of Appendix D for description) of the microbial communities are shown in Figure 5.10. The Shannon index decreased under all experimental conditions after 7 days (Figure 5.10a). The Hutcheson t-test showed that the Shannon indices on day 7 were significantly smaller (7 to 17%) compared with day 0 under all the treatment conditions (D.4 and D.5 of Appendix D). The Simpson index also decreased slightly between 0.02 and 2% after 7 days (Figure 5.10b). The decrease in both Shannon and Simpson indices indicates a decrease in richness for a given pattern of evenness, and/or a decrease in evenness for a given richness (Colwell, 2009). However, the magnitude of variation in the diversity indices is different, as Simpson diversity is less sensitive to the richness and more sensitive to evenness than the Shannon diversity (Colwell, 2009; Hill et al, 2003). The decrease in overall diversity is further supported by the observed decrease in both Chao1 and ACE richness estimators, as shown in Figures 5.10c and 5.10d. Overall, the approach used in this study involving inhibition agents to stimulate the growth of autotrophic and heterotrophic bacteria resulted in a slight decrease in microbial diversity.  00.751.5day 0HETHET+AOB+NOBAOB+NOBNOBday 7Relative abundance of known nitrifiers at genus level (%)Nitrosomonadaceae_unclassifiedNitrosomonasNitrosovibrioNitrospira90    Figure 5.10 Microbial diversity using a) Shannon index, b) Simpson index; microbial richness using c) Chao1 estimator, d) Abundance-Richness Coverage Estimator (ACE) in live-spiked and live-control microcosms. (HET: heterotrophic, AOB: ammonia-oxidizing bacteria, NOB: nitrite-oxidizing bacteria) Comparison between live-spiked and corresponding live-control microcosms revealed statistically significant differences in the Shannon diversity index for all treatment conditions, except HET (D.5 of Appendix D), with the spiked microcosm having lower diversity. Previous studies have reported a microbial community shift and decrease in diversity in river sediment and soil, resulting from exposure to a higher concentration of perfluoroalkyl acids (e.g., PFCAs) (Sun et al., 2016) and 6:2 FTS transformation products (Qiao et al., 2018; Zhang et al., 2017).   Effect of microbial activity on 6:2 FTS biotransformation The concentrations of 6:2 FTS in the live-spiked microcosms (Figure 5.11) show a decreasing trend, with about 7 and 20% of the initially spiked 6:2 FTS being biotransformed after 10 d (Figure 5.11). In contrast, the sterile control microcosms (Figure 5.11) showed no specific trend. The percentage of 6:2 FTS remaining after 10 d did not show a statistically significant difference under 91  various conditions (D.6 of Appendix D). However, Figure 5.11 reveals that in terms of 6:2 FTS biotransformation, the microcosms performed in the order of HET+AOB+NOB > AOB+NOB > NOB ≈ 1C HET. Overall higher 6:2 FTS biotransformation in the presence of ammonia-oxidation could be due to the presence of ammonium monooxygenase (AMO). AMO is a membrane-bound enzyme of AOB, which catalyzes the hydroxylation of ammonia to hydroxylamine, which is further oxidized to nitrite by hydroxylamine oxidoreductase (HAO). Monooxygenase enzymes have broad substrate range and low specificity (Fernandez-Fontaina et al., 2016; Tran et al., 2013). Desulfonation, considered to be the rate-limiting step in 6:2 FTS biotransformation (Wang et al., 2011), can be catalyzed by monooxygenases. A review of microbial desulfonation by Cook et al., (1998) showed that monooxygenase enzymes can destabilize C-SO3- bonds by inserting an oxygen atom to the same carbon, resulting in spontaneous loss of the sulfite group. In addition to copper-dependent monooxygenases (i.e., AMO), monooxygenases with other co-factors (e.g., reduced flavin mononucleotide (FMNH2), nicotinamide adenine dinucleotide phosphate (NADP)) are also known to increase the biotransformation of fluorotelomer compounds (Lewis et al., 2016; Van Hamme et al., 2013). Biological systems with nitrification are favored for on-site leachate treatment or off-site co-treatment with domestic wastewater (Berge et al., 2005) to remove the high ammonia content of landfill leachate. This study shows that the incorporation of nitrification in landfill leachate treatment schemes could significantly increase 6:2 FTS biotransformation.  92   Figure 5.11 Concentration of 6:2 FTS in live-spiked and sterile microcosms at various sampling days, normalized to initial spiking concentration (day 0). The initial spiking concentration varied between 750 – 800 µg/L (1756 – 1870 nmol/L). The absolute difference of duplicate measurements is represented with the error bars. (HET: heterotrophic, AOB: ammonia-oxidizing bacteria, NOB: nitrite-oxidizing bacteria).   Effect of microbial activity on PFCA formation  5.3.4.1 Formation of C4 to C6 PFCAs Following the biotransformation of 6:2 FTS, short-chain PFCAs (i.e., PFBA, PFPeA and PFHxA) increased in the live-spiked microcosms over the experimental period (Figure 5.12a). ΣC4 – C6 PFCAs after 10 d was higher under HET+AOB+NOB condition (~23 nmol/L) compared to HET (~8 nmol/L), AOB+NOB (~9 nmol/L) and NOB (~7 nmol/L) microsomes. A comparison between AOB+NOB and NOB also revealed slightly higher PFCAs formation in the presence of ammonia-oxidation. Previously, higher PFCAs formation in the presence of ammonia-oxidation in activated sludge has been attributed to the presence of AMO enzyme (Yu et al., 2018; Yu et al., 2016). Since 6:2 FTS biotransformation is cometabolic in nature (Yu et al., 2018b; Lewis et al., 2016; Kim et al., 2014), it is inherently connected to the presence of substrate. An overall greater amount of substrate (glucose+ammonia) for HET+AOB+NOB condition is likely to result in more complete 4060801001200 1 2 4 7 10mol % of 6:2 FTS spiked at day 0Time (days)HET AOB+NOBNOB HET+AOB+NOBSterile93  biotransformation (i.e., stable product formation). Observed ΣC4 – C6 PFCAs at 10 d, accounted for ~2 mol% of spiked 6:2 FTS at day 0 under HET+AOB+NOB condition. As discussed in section 5.3.2.2, a significantly higher abundance of Actinobacteria was observed under HET+AOB+NOB conditions (Figure 5.7), indicating that bacteria from this phylum could be resistant to 6:2 FTS and its biotransformation products and contribute towards the observed biotransformation and PFCA formation. Higher biotransformation products of 6:2 FTS observed under the HET+AOB+NOB condition could have led to the observed decrease in microbial diversity after 7 days when compared with HET+AOB+NOB condition without any spike as discussed in section 5.3.2.4. Furthermore, microbial species that grew under the HET+AOB+NOB condition could also be more sensitive towards 6:2 FTS biotransformation products. A recent study (Cai et al., 2019) found that for the same PFCA, growths of Escherichia coli and Pseudomonas putida are more affected than Arthrobacter strain GQ-9. Overall, this study suggests that exposure to short-chain PFCAs and possibly other 6:2 FTS biotransformation intermediates (e.g., fluorotelomer acids) can cause a shift in the microbial community in landfill at environmentally relevant concentrations (µg/L). However, long-term studies are needed to understand the implications of microbial community change resulting from exposure to short-chain PFCAs during biological leachate treatment on-site or off-site in WWTPs. About 18% of the transformation products were unaccounted for the HET+AOB+NOB condition, suggesting that major intermediates (e.g., saturated and unsaturated fluorotelomer acids, secondary fluorotelomer alcohols) were formed during 6:2 FTS biotransformation. Previous studies (Rand et al., 2014; Phillips et al., 2010, 2007) have demonstrated that metabolic intermediates of fluorotelomer compounds (e.g., saturated and unsaturated fluorotelomer acids) have higher cellular toxicity than PFCAs. Therefore, any regulatory and research efforts focused on exposure and toxicological risk assessment of PFASs from landfills should include fluorotelomer intermediates, in addition to PFCAs. The aerobic biotransformation of 6:2 FTS in landfill leachate-sediment, discussed in Chapter 3 (section 3.3), can provide valuable information with regards to possible fluorotelomer intermediates, their yield and transformation pathways.  94   Figure 5.12 Formation of a) C4 – C6 Perfluorocarboxylic acids (PFCAs) resulting from biotransformation of 6:2 FTS, and b) C7 to C8 PFCAs resulting from background biotransformation of unknown precursor compound in the live-spiked microcosms with heterotrophic (HET), ammonia-oxidizing bacteria (AOB) nitrite-oxidizing bacteria (NOB); The initial spiking concentration of 6:2 FTS varied between 750 - 800 µg/L in the live-spiked microcosms. The absolute differences of duplicate measurements are represented with error bars.  PFCA containing odd-numbered carbon (e.g., PFPeA) was only detected for HET+AOB+NOB conditions (Figure 5.12a). Previous studies have reported that the pathway and end products of 6:2 FTOH biotransformation are affected by microbial strain types, enzyme inducers and reducing energy (Kim et al., 2014; Kim et al., 2012). Therefore, it is likely that the microbial communities (discussed in section 5.3.2.2) and substrates added in the HET+AOB+NOB condition determined the rate of 6:2 FTS biotransformation and the products formed. 95  5.3.4.2 Background biotransformation In addition to C4 – C6 PFCAs, perfluoroheptanoic acid (PFHpA) and perfluorooctanoic acid (PFOA) also showed a modest increase in the live-spiked microcosms over the experimental period (Figure 5.12b). After 10 days, the sum of PFOA and PFHpA concentrations in live-controls (without 6:2 FTS spike) were comparable (~9 nmol/L) to that of the live spiked (Figure 5.12b). This suggests that background biotransformation of longer chain (≥C8) fluorotelomer precursor compounds, initially present in the landfill sediment used for inocula preparation, were responsible for the observed PFHpA and PFOA formation. For example, PFOA and PFHpA are biotransformation products of 8:2 FTOH in activated sludge and soil (Yu et al., 2016; Liu et al., 2007; Wang et al., 2005). Similar to previous discussions in section 5.3.4.1, the overall higher formation of PFOA and PFHpA was observed under HET+NOB+AOB conditions (Figure 5.12b). Similar to the previously observed occurrence of odd-chained PFCA (section 5.3.4.1, Figure 5.12a), PFHpA was also detected under HET+NOB+AOB conditions. 5.4 Conclusions and future research Aerobic biotransformation of 6:2 FTS and production of PFCAs were evaluated under the stimulated growth of heterotrophic (HET), ammonia-oxidizing (AOB) and nitrite-oxidizing bacteria (NOB) using inocula, prepared from sediment sampled from a leachate collection ditch. Both heterotrophic and autotrophic bacteria were able to biotransform 6:2 FTS to varying extents. Overall, greater biotransformation of 6:2 FTS observed in the presence of AOB likely resulted from the presence of ammonia monooxygenase (AMO) enzyme. It is recommended that future studies measuring AMO enzyme are needed to explore its effect on 6:2 FTS biotransformation. Higher PFCA formation (C4 – C6) was also observed under HET+AOB+NOB condition, possibly due to the overall higher amount of substrate (glucose+ammonia) added for this condition. Therefore, landfill leachate treatments with biological nitrification systems are likely to increase 6:2 FTS biotransformation and PFCA production.  Greater than 20-fold higher abundance of Actinobacteria was observed under the HET+AOB+NOB condition on day 7. As Actinobacteria can synthesize a wide range of enzymes including monooxygenases, they are likely to play an important role in 6:2 FTS biotransformation 96  and PFCA production. Microbial diversity (Shannon index) of spiked microcosms decreased in comparison with controls microcosms without the 6:2 FTS spike. Future research is needed to elucidate how the loss of microbial diversity might affect the performance of biological systems (e.g., activated sludge system, aerobic granular sludge) treating landfill leachate. The results from this study suggest that 6:2 FTS released with landfill leachate is a secondary source of short-chain PFCAs in the environment. Considering the mobile and persistent nature of short-chain PFCAs, their release would result in long-term exposure to humans, animals and biota. Short-chain PFCAs constituted a small fraction (<2%) of the parent compound, indicating that other fluorotelomer stable products and biotransformation intermediates (e.g., saturated and unsaturated fluorotelomer acids) would also be released to the aquatic environment.      97  Chapter 6: Phototransformation of 6:2 Fluorotelomer Sulfonate in Landfill Leachate Under Simulated Sunlight  6.1 Introduction 6:2 Fluorotelomer sulfonates (6:2 FTS) are widely used in surface coatings, paints, adhesives, wax and polishes as surfactants and water repellents (DuPont, 2012), applied to consumer products made of paper, wood, leather and textiles. In addition, 6:2 FTS is also used as a replacement of long-chain perfluorinated compound (e.g., perfluorooctane sulfonate (PFOS)) as a mist suppressant in chrome plating (Yang et al., 2014; Poulsen et al., 2011). Due to the disposal of consumer products containing 6:2 FTS at landfills, landfill leachates act as a secondary source of 6:2 FTS and its transformation products (e.g., C4 – C6 perfluorocarboxylic acids (PFCAs)) (Knutsen et al., 2019; Field and Seow 2017; Lang et al., 2017; Allred et al., 2014) in groundwater and surface water (Hepburn et al., 2019; Gobelius et al., 2018). Short-chain (C4 – C6) PFCAs are increasingly reported to dominate the poly- and perfluoroalkyl substance (PFAS) content in landfill leachate (Knutsen et al., 2019; Fuertes et al., 2017; Yan et al., 2015; Busch et al., 2010). Concerns exist regarding the high mobility and persistent nature of short-chain PFCAs, making them challenging to remove from drinking water (Li et al., 2019).  Previous research (Chapters 3 and 4) have shown that 6:2 FTS undergoes slow biotransformation (half-life >> 30 days) in landfill leachate under aerobic conditions. Due to the abstractable-H atoms in the fluorotelomer chain (see Figure E.1 of Appendix E), environmental oxidation processes (Butt et al., 2014) such as indirect phototransformation have the potential to affect the fate and transformation of 6:2 FTS in leachate, in addition to biotransformation. On-site leachate treatment options such as evaporation ponds, aerated lagoons, as well as off-site treatment in wastewater treatment plants (WWTPs) expose the leachates to sunlight. Following treatment, landfill leachate is often disposed of in surface waters (Townsend et al., 2015), thereby, making phototransformation a possible pathway for 6:2 FTS transformation in sunlit surface water. Previous studies (Trouborst, 2016; Gauthier and Mabury, 2005) have reported phototransformation of fluorotelomer compounds (i.e., 6:2 fluorotelomer sulfonamide 98  alkylbetaine (6:2 FTAB) and 8:2 fluorotelomer alcohol (8:2 FTOH)) in aqueous environments, with half-lives of 1 to 14 days (d), primarily producing persistent PFCAs (Trouborst, 2016; Gauthier and Mabury, 2005). Prior investigation of the phototransformation of 6:2 FTS in Milli-Q water using high energy far-UV radiation in the presence of a photosensitizer (e.g., H2O2) has reported short-chain (C2 to C7) PFCAs to be primary products (Yang et al., 2014). However, the aqueous phototransformation of 6:2 FTS under environmentally relevant conditions has not been studied to date. Aqueous phototransformation under solar irradiation consists of both direct and indirect phototransformation. Direct phototransformation occurs when the analyte of interest itself absorbs light radiation, resulting in a chemical reaction (Remucal, 2014). On the other hand, indirect phototransformation occurs when light radiations absorbed by photosensitive species, chemically produces reactive species, which then react with the analyte of interest (Schwarzenbach et al., 2003). Ubiquitous natural water and wastewater components (e.g., dissolved organic carbon (DOC), nitrate) absorbs light and reacts with dissolved oxygen to form reactive oxygen species (ROS), such as singlet oxygen (1O2), hydroxyl radicals (•OH), superoxide anions (O2•-), and H2O2 (Zhang et al., 2014) under solar radiation (as shown in Equations 6.1 and 6.2). In addition, the ROS can further react with natural water constituents (e.g., nitrate), forming reactive nitrogen species (Scholes et al., 2019). Recent studies have shown that, depending on their concentration and reactivity towards specific compounds, DOC (Ren et al., 2017; Wang et al., 2012; Xu et al., 2011) and nitrate (Scholes et al., 2019; Bonvin et al., 2013; Gauthier and Mabury, 2005) can play important roles in indirect phototransformation of organic pollutants, including fluorotelomer compounds (Trouborst, 2016; Gauthier and Mabury, 2005), in sunlit natural waters and engineered systems (e.g., aerated lagoons, constructed wetlands for leachate and wastewater treatment).     (6.1)     (6.2) ƕ 1O.OH DOC  DOCNO3- → NO2- +  .O- → .OH ƕ H+ 99  Landfill leachate is complex in nature, often characterized by high DOC and ammonia (tens to thousands of mg/L) (Townsend et al., 2015; Renou et al., 2008; Kjeldsen et al., 2002). Under aerobic conditions, the ammonia is converted to nitrate within days as discussed in section 4.3.1 of Chapter 4. Therefore, the environmental fate of 6:2 FTS in landfill leachate is likely affected by the DOC and nitrate contents of leachate, the extents of which are unknown at this time. The overall goal of this research project is to investigate the phototransformation of 6:2 FTS in landfill leachate under simulated sunlight conditions. The specific objectives included investigating the effects nitrate and humic acid (HA) on 6:2 FTS phototransformation, measuring the phototransformation rate of 6:2 FTS in landfill leachate and quantification of known transformation products (i.e., PFCAs) when subjected to simulated sunlight. 6.2 Materials and methods  Landfill leachate collection Landfill leachate was collected from a leachate collection sump in a municipal landfill (location is kept confidential at the request of the operator) using an ISCO 6712 portable sampling pump fitted with food-grade vinyl tubing. The tubing was flushed for ~2 minutes before collecting the leachate into 2 L polypropylene (PP) bottles. The samples were then placed with ice packs and transported to UBC within 2 h of collection. Following transport, the leachate was filtered through 0. 45 µm filter paper the same day and stored at -17°C. Prior to the use of the stored leachate for photodegradation experiments, physical and chemical characterization was performed as described in section 6.2.4.2.   Standards and reagents 6:2 FTS and four of its known phototransformation products, perfluorobutanoic acid (PFBA), perfluoropentanoic acid (PFPeA) and perfluorohexanoic acid (PFHxA), perfluoroheptanoic acid (PFHpA) (Yang et al., 2014), were monitored in all samples. Perfluorooctanoic acid (PFOA) and 5:3 fluorotelomer carboxylic acid (FTCA), compounds that are routinely detected in landfill leachates (Knutsen et al., 2019; Lang et al., 2017; Allred et al., 2014), were monitored in leachate samples. The analytes of interests (i.e., 6:2 FTS, PFBA, PFPeA, PFHxA, PFHpA, PFOA and 5:3 FTCA) were purchased from Wellington Laboratories, Guelph, Canada with >99% purity. The 100  acronyms and CAS numbers of the analytes of interest are listed in E.1 of Appendix E. HPLC grade acetonitrile and ammonium acetate were purchased from Sigma-Aldrich and Fisher Scientific, respectively. For all purposes, when water was needed, ultrapure deionized water (DI) from a Millipore® unit was used.  A phosphate buffer was selected to conduct the phototransformation experiments due to their high buffering capacity and high solubility in water. Potassium phosphate monobasic (KH2PO4) and dipotassium hydrogen orthophosphate (K2HPO4) were used to prepare the phosphate buffer solution (0.1 M) with neutral pH, the selected ionic strength of the phosphate buffer solution was within the previously reported range of 0.05 to 0.3 M for landfill leachate (Bradshaw et al., 2014; Guyonnet et al., 2005). A stock solution of nitrate (1000 mg/L) was prepared in DI. HA stock was prepared in 0.1 M potassium phosphate buffer (pH 7.1), to avoid the formation of precipitates. The suppliers of chemicals used for phosphate buffer and stock solution preparation are listed in E.2 of Appendix E. The actual concentrations of the prepared HA and nitrate stocks were 246 ± 1 and 984 ± 6 mg/L.   Experimental setup and sample collection 6.2.3.1 Experimental conditions Landfill leachate and 0.1 M phosphate buffer (pH 7.1) with HA and nitrate at various concentrations (Table 6.1), were used for the irradiation experiments. The irradiation solutions (16 mL) were prepared in 40 mL clear borosilicate vials (with PP caps) by spiking the phosphate buffer solution with respective stocks of HA and nitrate. Irradiation solutions were also spiked with 6:2 FTS diluted stock in DI (~100 mg/L) to achieve a concentration of ~100 µg/L (233 nm/L).      101  Table 6.1 Experimental conditions of 6:2 FTS phototransformation under simulated sunlighta Description Dissolved organic carbon (mg/L) Nitrate (mg/L) Humic acid (HA) in buffer solutionb 7 0 22 0 66 0 200 0 Humic acid-dark controlb 22 0 Nitrate in buffer solutionb 0 10 0 25 0 62 0 154 Nitrate-dark controlb 0 10 Humic acid and nitrate in buffer solutionb 22 25 pH 7.1 bufferb 0 0 pH 7.1 buffer-dark controlb 0 0 Deionized water (DI) 0 0 Leachate spiked 178 14 Leachate-dark control 178 14 Leachate-background 178 14 a all conditions were spiked with ∽100 µg/L of 6:2 FTS, except the leachate-background bpH 7.1 potassium phosphate buffer solution  6.2.3.2 Light soaking chamber setup The 6:2 FTS spiked solutions were irradiated for 72 hours (h) inside a light soaking chamber (Figure 6.1) in a lab in the UBC Chemistry department. Due to limited access to the lab, irradiation experiments lasting longer than 72 h were not feasible. The light soaking chamber was equipped with a 1000 W metal halide grow lamp of wavelength range 390 to 750 nm (Figure E.2 of Appendix E). The output of the lamp was set to 400 W dimmable electronic ballast. The samples were placed on a 15 cm × 20 cm area in a platform, 14 cm directly below the lamp. The output of the lamp was measured at the beginning and end of each experimental run, using a Hydrofarm LGBQM Quantum Photosynthetically Active Radiation (PAR) Meter (Hydrofarm, CA, USA). The measured PAR value varied between 850 and 1000 µmol/s/m2 on the platform area used for sample irradiation.  102   Figure 6.1 Schematic representation of the experimental setup showing the light source and sample arrangement. (See Figure E.3 for photo of light soaking chamber.) The temperature of the light soaking chamber was maintained constant at 39 ± 2°C (Figure E.4) using a built-in fan. The inside temperature of the chamber was monitored by continuous thermocouple measurements and a temperature input device (USB-TC01 from National Instruments™, USA), which included built-in software for viewing and logging data in a computer. Dark controls covered with aluminum foil were also run concurrently in a similar manner as the irradiated samples. At each sampling event (i.e., 0, 2, 4, 8, 24, 48, 72 h), 1 mL of irradiated and dark control samples were collected. This resulted in 7 mL out of the initial 16 mL solution being collected and stored at -17°C for further chemical analysis. Sample bottles were weighed after sample collection and before the next sampling time to calculate the evaporation losses and topped up with DI to correct for the losses. The pH of the samples was measured at 0 and 72 h. They did not change from 7.1 for the phosphate buffer solutions with HA and nitrate. The initial pH of 8.5 decreased slightly to ∽8.4 after 72 h for the landfill leachate samples.   Instrumental analysis 6.2.4.1 PFAS analysis To quantify the PFAS compounds, collected samples were spiked with M2_6:2FTS, M3PFBA, MPFHxA as the internal standards. In addition to the three labelled internal standards, leachate samples were spiked with M8PFOA. After vortex mixing, the samples were directly analyzed using HPLC-MS/MS. An Agilent 1200 series HPLC system (Agilent Technologies, CA, USA) 103  was used to separate the analytes in accordance with the instrumental method described in Chapter 3 section 3.2.5. Briefly, 4 µL of the sample was injected onto a Waters Xterra MS C18 column (100×2.1 mm, 3.5 µm particle size; Waters Corporation, Milford, MA), preceded by an Xterra MS C18 VanGuard cartridge (30×2.1 mm, 3.5 µm particle size Waters Corporation, Milford, MA). The mobile phase consisted of (A) water with 10 mM ammonium acetate and 10 mM acetic acid and (B) acetonitrile. The gradient profile of the mobile phases and detailed instrumental conditions are specified in E.3 of Appendix E.  6.2.4.2 Organic, inorganic and optical properties analysis A Shimadzu TOC analyzer (TOC-LCSH/CPN) was used for analyzing the DOC concentrations of the HA solutions and leachate samples. Anions were analyzed using a Dionex™ Aquion™ ion chromatography (IC) (Dionex Inc., Sunnyvale, CA, USA) system. Anions in IC were separated by an IonPac™ AS22 (4 × 250 mm) analytical column and IonPac™ AS22 guard column. The mobile phase consisted of 4.5 mM sodium carbonate and 1.4 mM sodium bicarbonate with a flow rate of 1.2 mL/min. UV/Vis transmittance was measured with a Thermo Spectronic Unicam UV 300 Spectrometer. For leachate and 1% 6:2 FTS aqueous solution, the measurements (shown in Figure 6.2) were done against a DI reference using 1 cm quartz cuvettes. The specific ultraviolet absorbance (SUVA254) was determined for the landfill leachate as the ratio of absorbance at 254 nm to the concentration of dissolved organic carbon. E2:E3 was calculated as the ratio of absorbance at 250 nm to that at 365 nm.   Quality control and quality assurance Throughout the irradiation experiment, sample processing and storage, polypropylene and HDPE were preferred over polytetrafluoroethylene (PTFE) containing materials to eliminate contamination by background fluorotelomer compounds. All irradiation experiments were conducted in at least two replicates. During LC-MS/MS analysis, quality control (QC) intermediate level calibration standards and instrumental blanks were injected every 10 – 15 samples to monitor absolute analyte areas, chromatographic retention times and background contamination from the instrument. The analytes of interest were below detection limits in all instrumental blanks in LC-MS/MS. The concentration of the QC standard varied within ± 15% of 104  the actual concentration. Instrumental analyses using the IC and TOC analyzers also included QC and blank samples to monitor absolute analyte areas, chromatographic retention times and background contamination. 6.3 Results and discussions  Physical and chemical and optical properties of landfill leachate The physical and chemical characterization and optical properties of the leachate are shown in Table 6.2. The pH of >8.5 suggested that the landfill is predominantly going through the methanogenic phase (Renou et al., 2008). The leachate collection system in the landfill studied also collects all the surface runoff from the landfill site. Therefore, the sampled leachate can be diluted by precipitation. This is supported by the observation that the DOC, ammonia, anion and cation concentrations of the leachate were closer to the lower bound of typical literature values (hundreds to thousands of mg/L for DOC, ammonia, anions and tens to hundreds of mg/L for cations as compiled in Table 3.2 in Chapter 3) (Kjeldsen et al., 2002) than the higher bound. Table 6.2 Characteristics of filtered (0.45 µm cellulose filter) landfill leachate (n=3) used in phototransformation experiments Characteristics  pH 8.6 ± 0.1% Electrical conductivity (mS/m) 25 ± 2% Alkalinity (mg CaCO3 mg/L) 1750 ± 2.5% Dissolved organic carbon (mg/L) 178 ± 1% Ammonia 134 ± 3% Nitrate (mg/L) 14 ± 0.7% Metals (mg/L)  Calcium  135 ± 5% Magnesium  53 ± 7% Silicon  28 ± 4% Anions (mg/L)  Sulfate 32 ± 2% Chloride 453 ± 2% Bromide  7.4 ± 2% Optical properties  Specific ultraviolet absorbance (SUVA254) (L/mg.m) 1.8 ± 1.3% E2:E3 (ratio of absorbance at 250 nm to that at 365 nm) 5.6 ± 1.8%  105  UV/Vis absorbance spectra (Figure 6.2 and optical properties in Table 6.2) of the leachate sample can provide insight into the structure of DOC (Maizel, 2017). The specific UV absorbance (SUVA254), a useful estimator of the aromaticity and molecular weight of DOC (Chowdhury, 2013; Hansen et al., 2016), was found to be ∽1.8 L /mg.m (Table 6.2) in the landfill leachate. This observed value was lower than the SUVA254 of DOC from terrestrially-dominated aquatic systems and the Sigma-Aldrich HA (typical range: 4 – 6 L /mg.m) (Al-Reasi et al., 2013; Weishaar et al., 2008), but similar to microbially-dominated aquatic systems (e.g., effluent from WWTP) (Maizel and Remucal, 2017). This indicates the presence of lower molecular weight compounds in leachate DOC, which absorbs less light, but can produce reactive species more efficiently compared with DOC of higher molecular weight (Maizel, 2017; MacKay et al., 2016; Helms et al., 2008).  Figure 6.2 Percentage transmittance of UV/Vis spectrum in filtered landfill leachate and 1% 6:2 FTS solution.  Effect of nitrate and humic acid on 6:2 FTS phototransformation The concentrations of spiked 6:2 FTS in phosphate buffer solutions containing nitrate and HA are shown in Figure 6.3. Over the 72 h experimental period, 6:2 FTS did not show any consistent decrease in the irradiated samples for the selected concentration ranges of nitrate and HA (Figures 6.3a to c). In addition, the concentration trends of the irradiated samples were similar to those of their respective dark controls (6.3a and b). This suggests little or no observable phototransformation of 6:2 FTS in the presence of nitrate and HA under the experimental 020406080100120260 310 360 410 460 510 560% of TransmittenceWavelength (nm)Landfill leachate106  conditions during 72 h of irradiation. Known phototransformation products of 6:2 FTS, C4 – C7 PFACs, were not detected in any of the samples. A previous study (Gauthier and Mabury, 2005) had shown that hydroxyl radical is primarily responsible for indirect phototransformation of fluorotelomer alcohol (8:2 FTOH). Therefore, the low yield of hydroxyl radicals under our experimental conditions would explain the observed lack of 6:2 FTS transformation. In addition, recent studies (Chen, 2019; Khosravifarsani et al., 2016) have shown that phosphate ions in the buffer have the capability to scavenge hydroxyl radicals, although to a lesser extent than common scavengers (e.g., tert-butyl alcohol) (Chen, 2019).  Phototransformation of 6:2 FTS in landfill leachate and product formation 6.3.3.1 Phototransformation of 6:2 FTS  The spiked 6:2 FTS showed a consistent decreasing trend in irradiated landfill leachate as opposed to the leachate dark control sample (Figure 6.4), suggesting that 6:2 FTS can undergo phototransformation in landfill leachate in a sunlight drenched environment. However, the absence of a decrease in spiked 6:2 FTS in irradiated DI suggests that direct phototransformation was not occurring, which can be attributed to the lack of absorbance of 6:2 FTS at environmentally relevant wavelengths (λ>290 nm) (Figure 6.2). Therefore, indirect phototransformation was likely responsible for the observed decrease of 6:2 FTS in landfill leachate. The phototransformation of 6:2 FTS in landfill leachate followed pseudo-first-order kinetics, with a rate constant of 0.0039 h-1. The estimated half-life of 6:2 FTS was ∽178 h (coefficient of determination, R2>0.94 for 7 data points as shown in Figure E.5), roughly equivalent to ∽15 d considering the day and night cycle. Previous studies (Trouborst, 2016; Gauthier and Mabury, 2005) have reported half-lives ranging between 30 and 163 h, and 14 and 108 h for 8:2 FTOH and 6:2 FTAB in lake water and synthetic field water containing DOC, nitrate and bicarbonate, under simulated sunlight. The slightly higher half-life of 6:2 FTS observed in leachate in our study compared to previous studies (Trouborst, 2016; Gauthier and Mabury, 2005) could be due to differences in irradiation matrix and light source.   107   Figure 6.3 Time variation of spiked 6:2 FTS in pH 7.1 phosphate buffer solutions in the presence of a) 25 – 154 mg/ L nitrate, b) 7 – 200 mg/L of humic acid (HA), and c) with 22 and 25 mg/L HA and nitrate. The initial spiking concentration of 6:2 FTS ranged between 110 – 130 µg/L. The errors of duplicate measurements (or triplicates for 10 mg/L nitrate) are represented by the error bars. The errors were within the range of those observed for quality control standards (±15%). 108  While the phosphate buffer solutions with nitrate and HA did not show measurable transformation (section 6.3.2), 6:2 FTS decreased in landfill leachate (Figure 6.4) with nitrate and DOC content comparable to that of the phosphate buffer solutions. Previous studies (Sharpless, 2012; Halladja et al., 2007) have shown that the photoreactivity of DOC is closely related to the structure and source of the DOC. Therefore, differences in the structure of DOC leachate compared to HA solution could contribute to increased formation of reactive species, resulting in 6:2 FTS transformation. The landfill leachate DOC has been shown to include hydrophilic, humic and fulvic acid fractions (Xie and Guan, 2014; Driskill, 2013; Zhang et al., 2013; Huo et al., 2008). As discussed in section 3.1, the SUVA254 of leachate indicated the presence of lower molecular weight DOCs, which are capable of efficient production of reactive species (Maizel, 2017; MacKay et al., 2016; Helms et al., 2008).   Figure 6.4 Phototransformation of spiked 6:2 FTS in landfill leachate in irradiated and dark control solution over a period of 72 hours. The initial spiking concentration was ∽100 µg/L. The absolute differences of duplicate (and triplicates for leachate spiked conditions) are expressed by the error bars.  6.3.3.2 Phototransformation products of 6:2 FTS Among the PFACs investigated, PFBA and PFHxA were above the detection limit in all leachate samples (Figure 6.5b). PFHpA and PFPeA were not detected in any of the leachate samples. The PFBA concentration remained almost constant throughout the experimental period in irradiated and control samples. The increase of PFHxA only in irradiated leachate spiked and background 0204060801001201400 10 20 30 40 50 60 70 80% of initially spiked 6:2 FTSTime (Hours)Leachate spikedLeachate spiked-dark controlDI water109  samples (without 6:2 FTS) suggests that the formation of PFHxA was likely a result of phototransformation of 6:2 FTS or other fluorotelomer precursors present in leachate. For example, 5:3 FTCA, a common biotransformation product of fluorotelomer compounds in leachate, was present in all leachate samples at concentrations ranging between 2 and 3 µg/L (Figure 6.5b). Lack of quantifiable PFCAs indicates the formation of unknown products and possible loss of volatile products (e.g., 6:2 fluorotelomer alcohol and aldehyde) that were not monitored in this study. For example, Yang et al., (2014) reported a novel intermediate (proposed structure: CF3(CF2)5CH(OH)CH(OH)SO3-) to be a major product of advanced photooxidation of 6:2 FTS. Another study (Gauthier and Mabury, 2005). reported that ∽2% of initially spiked 8:2 FTOH was converted to known products (e.g., FTCA, PFCA etc.) in synthetic field water containing DOC, nitrate and bicarbonate. For future studies, an irradiation time longer that 72 h is recommended, which would allow quantification of the phototransformation products and performance of a mass balance analysis.  110   Figure 6.5 a) C4 and C6 perfluorocarboxylic acids (PFCAs), and b) 5:3 fluorotelomer carboxylic acid (FTCA) in landfill leachate in irradiated and dark control solution over 72 hour period. The absolute differences of duplicate are expressed as error bars.  6.4 Conclusions and future research  Phototransformation of spiked 6:2 FTS was investigated in pH 7.1 phosphate buffer solution in the presence of nitrate and humic acid, and in landfill leachate under simulated sunlight. The spiked 6:2 FTS did not show any observable decrease in the presence of nitrate and humic acid over 72 h, likely due to the low yield of hydroxyl radicals. However, 6:2 FTS underwent indirect phototransformation in landfill leachate, with an estimated half-life of ∽15 d. This suggests that indirect phototransformation of 6:2 FTS is likely a relevant environmental transformation pathway in diluted leachate. The increase of PFHxA in irradiated leachate background samples (without 111  6:2 FTS) suggested that phototransformation in sunlit leachate could lead to formation of persistent PFCAs, even at environmental concentrations of the precursors. To better understand the mechanism of indirect phototransformation of 6:2 FTS, the roles of various reactive species need to be studied using specific probe compounds. In addition, to better elucidate the role of leachate DOC, phototransformation experiments should be conducted with various fractions of DOC isolated from landfill leachate.   112  Chapter 7: Conclusions and Recommendations for Future Work 7.1 Conclusions Fluorotelomer compounds (FTCs) consists of a major fraction of the important class of industrial chemicals called poly- and perfluoroalkyl substances (PFASs). Consumer products and packaging containing FTCs are disposed of at municipal landfills at the end of their useful lives, where they undergo abiotic and biotic transformations. While previous studies have routinely quantified FTCs (e.g., fluorotelomer alcohol (FTOH) and fluorotelomer sulfonate (FTS)) and their known biotransformation products (e.g., perfluorocarboxylic acid (PFCA)) in landfill leachate, their environmental fate and transformation in leachate have not been studied to date. Therefore, this study was focused on measuring the transformation of FTCs in landfill leachate using lab-scale experiments under environmentally relevant conditions. The findings from the aerobic biotransformation studies showed that 8:2 FTOH and 6:2 FTS persisted in landfill leachate and sediment (half-life >>30 d). Slower biotransformation could result in significant partitioning of 8:2 FTOH to the atmosphere during on-site leachate storage and treatment, especially involving aeration (e.g., aerated lagoons, air stripping, etc.). The biotransformation pathways of 8:2 FTOH and 6:2 FTS were similar to those observed in soil, sediment and activated sludge, with PFCAs being the primary end products. These suggest that landfills act as secondary sources of FTCs and their biotransformation products, including legacy pollutants such as perfluorooctanoic acid (PFOA) and highly mobile short-chain (C4 – C6) PFCAs in the environment. For the first time, this study reported that dilution of landfill leachate, (e.g., during wet seasons), would likely result in reduced 6:2 FTS biotransformation and increased PFCAs formation compared to dry seasons. 6:2 FTS biotransformation under stimulated growth of heterotrophic, ammonia-oxidizing and nitrite-oxidizing bacteria, showed that both heterotrophic and nitrifying bacteria were able to biotransform 6:2 FTS to varying extents. However, greater biotransformation of 6:2 FTS and higher PFCA formation (C4 – C6) were observed in microcosms with heterotrophic and nitrifier growths. This suggests that biological landfill leachate treatments with nitrification systems are likely to increase 6:2 FTS biotransformation and cause greater PFCA release with treated leachate. 6:2 FTS also underwent phototransformation in landfill leachate under simulated sunlight, with an estimated half-life of ∽15 days, indicating that indirect 113  phototransformation of 6:2 FTS is likely a relevant environmental transformation pathway in sunlit aquatic environments.  Overall, FTCs exhibited chain length- and functional-group-dependent transformation behavior in a landfill leachate-sediment system, producing a mixture of partially- and fully fluorinated compounds with distinct partitioning behavior and varying persistence. Therefore, strategies including monitoring of the waste, robust multi-media sampling and pollution control measures are needed to mitigate risks of PFAS compounds in landfill leachate. 7.2 Recommendations for future research Based on the observations and findings from this study, the following recommendations are made for future research. • The landfill leachate used in this study can be considered dilute (TOC < 1000 mg/L), compared to other municipal solid waste landfills. Also, the leachate collected here likely represents a mixture of leachate from various stages of waste decomposition. Therefore, further investigation is recommended to better elucidate the effect of leachate from various stages of waste decomposition and highly concentrated leachate on FTC biotransformation.  • Formation of the sediment-bound residue of FTCs observed here, can potentially reduce toxicity bioavailability of FTCs. Long-term stability and remobilization potential of such residues from sediment/soli contaminated with leachate management operations (e.g., evaporation ponds, leachate ditches, aerated lagoons) need to be studied.  • Biotransformation of FTCs primarily produced PFCAs in landfill leachate. Considering the persistent nature and high mobility of short-chain PFCAs, future research is necessary to develop effective removal techniques for PFCAs from landfill leachate to limit their release in the environment. • Measurement of key enzymes (i.e., ammonium monooxygesane) is recommended to confirm the positive effect of ammonia-oxidizers on FTC biotransformation during nitrification.  114  • Future studies are needed to elucidate the relationship between heterotrophic and autotrophic biomass growth and FTC biotransformation, to develop a cometabolic model that can predict the removal of parent compounds and PFCAs generation in leachate biological treatment systems.  • To better understand the mechanism of indirect phototransformation of 6:2 FTS, the role of various reactive species needs to be studied using specific probe compounds.  • To better elucidate the role of dissolved organic carbon on indirect phototransformation of FTCs, phototransformation experiments should be conducted with dissolved organic carbon isolated from landfill leachate. 115  References  Abed, R. M., Safi, N. M., Köster, J., De Beer, D., El-Nahhal, Y., Rullkötter, J., & Garcia-Pichel, F. (2002). Microbial diversity of a heavily polluted microbial mat and its community changes following degradation of petroleum compounds. Applied and Environmental Microbiology, 68(4), 1674-1683. https://doi.org/10.1128/AEM.68.4.1674-1683.2002 Ahrens, L. (2011). Polyfluoroalkyl compounds in the aquatic environment: a review of their occurrence and fate. Journal of Environmental Monitoring, 13(1), 20-31. https://doi.org/10.1039/C0EM00373E  Ahrens, L., Shoeib, M., Harner, T., Lee, S.C., Guo, R., Reiner, E.J. (2011). Wastewater treatment plant and landfills as sources of polyfluoroalkyl compounds to the atmosphere. Environmental Science & Technology 45, 8098-8105. https://doi.org/10.1021/es1036173. Allen, M. R., Braithwaite, A., & Hills, C. C. (1997). Trace organic compounds in landfill gas at seven UK waste disposal sites. Environmental Science & Technology, 31(4), 1054-1061. https://doi.org/10.1021/es9605634 Allred, B. M., Lang, J. R., Barlaz, M. A., & Field, J. A. (2014). Orthogonal zirconium diol/C18 liquid chromatography-tandem mass spectrometry analysis of poly and perfluoroalkyl substances in landfill leachate. Journal of Chromatography A, 1359, 202-211. https://doi.org/10.1016/j.chroma.2014.07.056 Allred, B.M., Lang, J.R., Barlaz, M.A., Field, J.A. (2015). Physical and biological release of poly- and perfluoroalkyl substances (PFASs) from municipal solid waste in anaerobic model landfill reactors. Environmental Science & Technology. 49, 7648-7656. https://doi.org/10.1021/acs.est.5b01040 Alvarez, A., Saez, J. M., Costa, J. S. D., Colin, V. L., Fuentes, M. S., Cuozzo, S. A., & Amoroso, M. J. (2017). Actinobacteria: current research and perspectives for bioremediation of pesticides and heavy metals. Chemosphere, 166, 41-62. http://dx.doi.org/10.1016/j.chemosphere.2016.09.070 Alvarez-Cohen, L., & Speitel, G. E. (2001). Kinetics of aerobic cometabolism of chlorinated solvents. Biodegradation, 12(2), 105-126. Arvaniti, O. S., & Stasinakis, A. S. (2015). Review on the occurrence, fate and removal of perfluorinated compounds during wastewater treatment. Science of the Total Environment, 524, 81-92. https://doi.org/10.1016/j.scitotenv.2015.04.023  Arvaniti, O.S., Asimakopoulos, A.G., Dasenaki, M.E., Ventouri, E.I., Stasinakis, A.S., Thomaidis, N.S. (2014). Simultaneous determination of eighteen perfluorinated compounds in dissolved and particulate phases of wastewater, and in sewage sludge by liquid chromatography-tandem mass spectrometry. Anal Methods. 6, 1341-1349. https://doi.org/10.1039/C3AY42015A. Arvaniti, O.S., Ventouri, E.I., Stasinakis, A.S., Thomaidis, N.S. (2012). Occurrence of different classes of perfluorinated compounds in Greek wastewater treatment plants and determination of their solid-water distribution coefficients. Journal of Hazardaous Materials. 239, 24-31. https://doi.org/10.1016/j.jhazmat.2012.02.015. 116  Ateia, M., Maroli, A., Tharayil, N., & Karanfil, T. (2019). The overlooked short-and ultrashort-chain poly-and perfluorinated substances: A review. Chemosphere, 220, 866-882. https://doi.org/10.1016/j.chemosphere.2018.12.186 Backe, W. J., Day, T. C., & Field, J. A. (2013). Zwitterionic, cationic, and anionic fluorinated chemicals in aqueous film forming foam formulations and groundwater from US military bases by nonaqueous large-volume injection HPLC-MS/MS. Environmental Science & Technology, 47(10), 5226-5234. https://doi.org/10.1021/es3034999 Banzhaf, S., Filipovic, M., Lewis, J., Sparrenbom, C. J., & Barthel, R. (2017). A review of contamination of surface-, ground-, and drinking water in Sweden by perfluoroalkyl and polyfluoroalkyl substances (PFASs). Ambio, 46(3), 335-346. https://doi.org/10.1007/s13280-016-0848-8 Barns, S. M., Cain, E. C., Sommerville, L., and Kuske, C. R. (2007). Acidobacteria phylum sequences in uranium-contaminated subsurface sediments greatly expand the known diversity within the phylum. Applied and Environmental Microbiology 73, 3113–3116. https://doi.org/10.1128/AEM.02012-06 Bečanová, J., Melymuk, L., Vojta, Š., Komprdová, K., & Klánová, J. (2016). Screening for perfluoroalkyl acids in consumer products, building materials and wastes. Chemosphere, 164, 322-329. https://doi.org/10.1016/j.chemosphere.2016.08.112  Benskin, J. P., Li, B., Ikonomou, M. G., Grace, J. R., & Li, L. Y. (2012). Per-and polyfluoroalkyl substances in landfill leachate: patterns, time trends, and sources. Environmental Science & Technology, 46(21), 11532-11540. https://doi.org/10.1021/es302471n Berge, N. D., Reinhart, D. R., & Townsend, T. G. (2005). The fate of nitrogen in bioreactor landfills. Critical Reviews in Environmental Science and Technology, 35(4), 365-399. https://doi.org/10.1080/10643380590945003 Bollag, J. M. (1991). Enzymatic binding of pesticide degradation products to soil organic matter and their possible release. In ACS Symposium series American Chemical Society. https://doi.org/10.1021/bk-1991-0459.ch009 Bonvin, F., Omlin, J., Rutler, R., Schweizer, W. B., Alaimo, P. J., Strathmann, T. J., & Kohn, T. (2013). Direct photolysis of human metabolites of the antibiotic sulfamethoxazole: evidence for abiotic back-transformation. Environmental Science & Technology, 47(13), 6746-6755. https://doi.org/10.1021/es303777k Bossi, R., Strand, J., Sortkjaer, O., Larsen, M.M. (2008). Perfluoroalkyl compounds in Danish wastewater treatment plants and aquatic environments. Environ Int. 34, 443-450. https://doi.org/10.1016/j.envint.2007.10.002. Bradshaw, S. L., & Benson, C. H. (2014). Effect of municipal solid waste leachate on hydraulic conductivity and exchange complex of geosynthetic clay liners. Journal of Geotechnical and Geoenvironmental Engineering, 140(4), 04013038. https://doi.org/10.1061/(ASCE) GT.1943-5606.0001050 Braker, G., & Conrad, R. (2011). Diversity, structure, and size of N2O-producing microbial communities in soils—what matters for their functioning? In Advances in applied 117  microbiology (Vol. 75, pp. 33-70). Academic Press. https://doi.org/10.1016/B978-0-12-387046-9.00002-5 Brendel, S., Fetter, É., Staude, C., Vierke, L., & Biegel-Engler, A. (2018). Short-chain perfluoroalkyl acids: environmental concerns and a regulatory strategy under REACH. Environmental Sciences Europe, 30(1), 9. https://doi.org/10.1186/s12302-018-0134-4 Buck, R. C., Franklin, J., Berger, U., Conder, J. M., Cousins, I. T., De Voogt, P., & van Leeuwen, S. P. (2011). Perfluoroalkyl and polyfluoroalkyl substances in the environment: terminology, classification, and origins. Integrated Environmental Assessment and Management, 7(4), 513-541. https://doi.org/10.1002/ieam.258 Busch, J., Ahrens, L., Sturm, R., & Ebinghaus, R. (2010). Polyfluoroalkyl compounds in landfill leachates. Environmental Pollution, 158(5), 1467-1471. https://doi.org/10.1016/j.envpol.2009.12.031 Butt, C. M., Muir, D. C., & Mabury, S. A. (2014). Biotransformation pathways of fluorotelomer‐based polyfluoroalkyl substances: A review. Environmental Toxicology and Chemistry, 33(2), 243-267. https://doi.org/10.1002/etc.2407 Cai, M., Zhao, Z., Yin, Z., Ahrens, L., Huang, P., Cai, M., & Xie, Z. (2011). Occurrence of perfluoroalkyl compounds in surface waters from the North Pacific to the Arctic Ocean. Environmental Science & Technology, 46(2), 661-668. https://doi.org/10.1021/es2026278 Cai, Y., Chen, H., Chen, H., Li, H., Yang, S., & Wang, F. (2019). Evaluation of Single and Joint Toxicity of Perfluorinated Carboxylic Acids and Copper to Metal-Resistant Arthrobacter Strains. International Journal Of Environmental Research And Public Health, 16(1), 135. https://doi.org/10.3390/ijerph16010135 Cederbaum, A. I. (2012). Alcohol metabolism. Clinics in liver disease, 16(4), 667-685. https://doi.org/10.1016/j.cld.2012.08.002 Chen, H. Y. (2019). Why the Reactive Oxygen Species of the Fenton Reaction Switches from Oxoiron (IV) Species to Hydroxyl Radical in Phosphate Buffer Solutions? A Computational Rationale. ACS Omega, 4(9), 14105-14113. https://doi.org/10.1021/acsomega.9b02023 Chowdhury, S. (2013). Trihalomethanes in drinking water: Effect of natural organic matter distribution. Water SA, 39(1), 1-8. https://doi.org/10.4314/wsa.v39i1.1 Clarke, B. O., Anumol, T., Barlaz, M., & Snyder, S. A. (2015). Investigating landfill leachate as a source of trace organic pollutants. Chemosphere, 127, 269-275. https://doi.org/10.1016/j.chemosphere.2015.02.030 Colwell, R. K. (2009). Biodiversity: concepts, patterns, and measurement. The Princeton guide to ecology, 257-263 Cook A.M., Laue H., Junker F. (1998). Microbial desulfonation, FEMS Microbiology Reviews, 22, 399–419. https://doi.org/10.1111/j.1574-6976.1998.tb00378.x Cottrell, M. T., & Kirchman, D. L. (2000). Natural assemblages of marine proteobacteria and members of the Cytophaga-Flavobacter cluster consuming low-and high-molecular-weight 118  dissolved organic matter. Applied and Environmental Microbiology, 66(4), 1692-1697. https://doi.org/10.1128/AEM.66.4.1692-1697.2000 Cousins, I. T., Vestergren, R., Wang, Z., Scheringer, M., & McLachlan, M. S. (2016). The precautionary principle and chemicals management: The example of perfluoroalkyl acids in groundwater. Environment International, 94, 331-340. https://doi.org/10.1016/j.envint.2016.04.044 D’eon, J. C., & Mabury, S. A. (2011). Is indirect exposure a significant contributor to the burden of perfluorinated acids observed in humans? Environmental Science & Technology, 45(19), 7974-7984. https://doi.org/10.1021/es200171y Dalton, H., & Stirling, D. I. (1982). Co-metabolism. Philosophical Transactions of the Royal Society of London. B, Biological Sciences, 297(1088), 481-496. https://doi.org/10.1098/rstb.1982.0056 De Silva, A. O., Allard, C. N., Spencer, C., Webster, G. M., & Shoeib, M. (2012). Phosphorus-containing fluorinated organics: polyfluoroalkyl phosphoric acid diesters (diPAPs), perfluorophosphonates (PFPAs), and perfluorophosphinates (PFPIAs) in residential indoor dust. Environmental Science & Technology, 46(22), 12575-12582. https://doi.org/10.1021/es303172p Dec, J., & Bollag, J. M. (1997). Determination of covalent and noncovalent binding interactions between xenobiotic chemicals and soil. Soil Science, 162(12), 858-874. Dinglasan, M. J. A., Ye, Y., Edwards, E. A., & Mabury, S. A. (2004). Fluorotelomer alcohol biodegradation yields poly-and perfluorinated acids. Environmental Science & Technology, 38(10), 2857-2864. https://doi.org/10.1021/es0350177 Dinglasan-Panlilio, M. J. A., & Mabury, S. A. (2006). Significant residual fluorinated alcohols present in various fluorinated materials. Environmental Science & Technology, 40(5), 1447-1453. https://doi.org/10.1021/es051619+ Donova, M. V. (2007). Transformation of steroids by actinobacteria: a review. Applied Biochemistry and Microbiology, 43(1), 1-14. https://doi.org/10.1134/S0003683807010012 Driskill, N. M. (2013). Characterization and treatment of organic matter, UV quenching substances, and organic nitrogen in landfill leachates. Masters dissertation, Virginia Polytechnic Institute and State University  Du, Z., Deng, S., Bei, Y., Huang, Q., Wang, B., Huang, J., & Yu, G. (2014). Adsorption behavior and mechanism of perfluorinated compounds on various adsorbents—A review. Journal of Hazardous Materials, 274, 443-454. https://doi.org/10.1016/j.jhazmat.2014.04.038 DuPont (2010). DuPont Surface Protection Solutions. Available at https://cms.chempoint.com/ChemPoint/media/ChemPointSiteMedia/PDF%20Docs/K-23574-Zonyl-to-Capstone-Transition-Guide-Brochure.pdf Accessed on June, 2019 DuPont (2012) DuPont Surface Protection Solutions: Dupont™ Capstone® Repellents And Surfactants. Available at https://cms.chempoint.com/ChemPoint/media/ChemPointSiteMedia/PDF%20Docs/K-119  20614-2-DuPont-Capstone-Product-Stewardship-Detail-Document.pdf accessed on June 2019. Durmusoglu, Ertan, Fatih Taspinar, and Aykan Karademir. "Health risk assessment of BTEX emissions in the landfill environment." Journal of Hazardous Materials 176.1-3 (2010): 870-877. https://doi.10.1016/j.jhazmat.2009.11.117 Eggen, T., Moeder, M., & Arukwe, A. (2010). Municipal landfill leachates: a significant source for new and emerging pollutants. Science of the Total Environment, 408(21), 5147-5157. https://doi.org/10.1016/j.scitotenv.2010.07.049 Ellis, R. J., Morgan, P., Weightman, A. J., and Fry, J. C. (2003). Cultivation dependent and -independent approaches for determining bacterial diversity in heavy-metal-contaminated soil. Applied and Environmental Microbiology 69, 3223–3230. https://doi.org/10.1128/AEM.69.6.3223-3230.2003 Fernandez-Fontaina, E., Gomes, I. B., Aga, D. S., Omil, F., Lema, J. M., & Carballa, M. (2016). Biotransformation of pharmaceuticals under nitrification, nitratation and heterotrophic conditions. Science of the Total Environment, 541, 1439-1447. https://doi.org/10.1016/j.scitotenv.2015.10.010 Fernández-Gomez, B., Richter, M., Schüler, M., Pinhassi, J., Acinas, S. G., González, J. M., & Pedros-Alio, C. (2013). Ecology of marine Bacteroidetes: a comparative genomics approach. The ISME journal, 7(5), 1026. https://doi.org/10.1038/ismej.2012 Field, J. A., & Seow, J. (2017). Properties, occurrence, and fate of fluorotelomer sulfonates. Critical Reviews in Environmental Science and Technology, 47(8), 643-691. https://doi.org/10.1080/10643389.2017.1326276 Filipovic, M., Laudon, H., McLachlan, M. S., & Berger, U. (2015). Mass balance of perfluorinated alkyl acids in a pristine boreal catchment. Environmental Science & Technology, 49(20), 12127-12135. https://doi.org/10.1021/acs.est.5b03403 Fischer, K., & Majewsky, M. (2014). Cometabolic degradation of organic wastewater micropollutants by activated sludge and sludge-inherent microorganisms. Applied Microbiology and Biotechnology, 98(15), 6583-6597. https://doi.org/10.1007/s00253-014-5826-0 Foo, K., Hameed, B., (2009). An overview of landfill leachate treatment via activated carbon adsorption process. Journal of Hazardous Materials, 171, 54-60. https://doi.org/10.1016/j.jhazmat.2009.06.038. Fuertes, I., Gómez-Lavín, S., Elizalde, M., Urtiaga, A., (2017). Perfluorinated alkyl substances (PFASs) in northern Spain municipal solid waste landfill leachates. Chemosphere. 168, 399-407.  https://doi.org/10.1016/j.chemosphere.2016.10.072. Gallen, C., D. Drage, G. Eaglesham, S. Grant, M. Bowman, and J. F. Mueller., (2017). Australia-wide assessment of perfluoroalkyl substances (PFASs) in landfill leachates. Journal of Hazardous Materials, 331, 132-141. https://doi.org/10.1016/j.jhazmat.2017.02.006 Gallen, C., Drage, D., Kaserzon, S., Baduel, C., Gallen, M., Banks, A., Broomhall, S., Mueller, J., (2016). Occurrence and distribution of brominated flame retardants and perfluoroalkyl 120  substances in Australian landfill leachate and biosolids. Journal of Hazardous Materials, 312, 55-64. https://doi.org/10.1016/j.jhazmat.2016.03.031 Gallen, C., Eaglesham, G., Drage, D., Nguyen, T. H., & Mueller, J. F. (2018). A mass estimate of perfluoroalkyl substance (PFAS) release from Australian wastewater treatment plants. Chemosphere, 208, 975-983. https://doi.org/10.1016/j.chemosphere.2018.06.024 Gao, J., Oloibiri, V., Chys, M., Audenaert, W., Decostere, B., He, Y., & Van Hulle, S. W. (2015). The present status of landfill leachate treatment and its development trend from a technological point of view. Reviews in Environmental Science and Biotechnology, 14(1), 93-122. https://doi.org/10.1007/s11157-014-9349-z Gardener, M. (2012). Statistics for ecologists using R and Excel. Data collection, exploration. Analysis and Presentation. Pelagic Publishing, Exeter. Gauthier, S. A., & Mabury, S. A. (2005). Aqueous photolysis of 8: 2 fluorotelomer alcohol. Environmental Toxicology and Chemistry, 24(8), 1837-1846. https://doi.org/10.1897/04-591R.1 Gianfreda, L., & Bollag, J. M. (2002). Isolated enzymes for the transformation and detoxification of organic pollutants (pp. 495r-538). Marcel Dekker: New York. Gobelius, L., Hedlund, J., Dürig, W., Tröger, R., Lilja, K., Wiberg, K., & Ahrens, L. (2018). Per-and polyfluoroalkyl substances in Swedish groundwater and surface water: implications for environmental quality standards and drinking water guidelines. Environmental Science & Technology, 52(7), 4340-4349. https://doi.org/10.1021/acs.est.7b05718 Gomis, M. I., Vestergren, R., MacLeod, M., Mueller, J. F., & Cousins, I. T. (2017). Historical human exposure to perfluoroalkyl acids in the United States and Australia reconstructed from biomonitoring data using population-based pharmacokinetic modelling. Environment International, 108, 92-102. https://doi.org/10.1016/j.envint.2017.08.002 Grady Jr, C. L., Daigger, G. T., Love, N. G., & Filipe, C. D. (2011). Biological Wastewater Treatment. CRC press. Greer, C. W., Whyte, L. G., & Niederberger, T. D. (2010). Microbial communities in hydrocarbon-contaminated temperate, tropical, alpine, and polar soils. Handbook Of Hydrocarbon And Lipid Microbiology, 2313-2328. https://doi.org/10.1007/978-3-540-77587-4_168 Guerra, P., Kim, M., Kinsman, L., Ng, T., Alaee, M., Smyth, S.A., (2014). Parameters affecting the formation of perfluoroalkyl acids during wastewater treatment. Journal of Hazardous Materials, 272, 148-154.  https://doi.org/10.1016/j.jhazmat.2014.03.016. Gupta, R. S. (2000). The phylogeny of proteobacteria: relationships to other eubacterial phyla and eukaryotes. FEMS Microbiology Reviews, 24(4), 367-402. https://doi.org/10.1111/j.1574-6976.2000.tb00547.x Guyonnet, D., Gaucher, E., Gaboriau, H., Pons, C. H., Clinard, C., Norotte, V., & Didier, G. (2005). Geosynthetic clay liner interaction with leachate: correlation between permeability, microstructure, and surface chemistry. Journal of Geotechnical and Geoenvironmental Engineering, 131(6), 740-749. https://doi.org/10.1061/~ASCE!1090-0241~2005!131:6~740 121  Halladja, S., Ter Halle, A., Aguer, J. P., Boulkamh, A., & Richard, C. (2007). Inhibition of humic substances mediated photooxygenation of furfuryl alcohol by 2, 4, 6-trimethylphenol. Evidence for reactivity of the phenol with humic triplet excited states. Environmental Science & Technology, 41(17), 6066-6073. https://doi.org/10.1021/es070656t Han, Y., Zhang, F., Wang, Q., Zheng, S., Guo, W., Feng, L., & Wang, G. (2016). Flavihumibacter stibioxidans sp. nov., an antimony-oxidizing bacterium isolated from antimony mine soil. International Journal of Systematic and Evolutionary Microbiology, 66(11), 4676-4680. https://doi.org/10.1099/ijsem.0.001409 Hansen, A. M., Kraus, T. E., Pellerin, B. A., Fleck, J. A., Downing, B. D., & Bergamaschi, B. A. (2016). Optical properties of dissolved organic matter (DOM): Effects of biological and photolytic degradation. Limnology and Oceanography, 61(3), 1015-1032. https://doi.org/10.1002/lno.10270 Harding-Marjanovic, K. C., Houtz, E. F., Yi, S., Field, J. A., Sedlak, D. L., & Alvarez-Cohen, L. (2015). Aerobic biotransformation of fluorotelomer thioether amido sulfonate (Lodyne) in AFFF-amended microcosms. Environmental Science & Technology, 49(13), 7666-7674. https://doi.org/10.1021/acs.est.5b01219 Harrad, S., Drage, D. S., Sharkey, M., & Berresheim, H. (2020). Perfluoroalkyl substances and brominated flame retardants in landfill-related air, soil, and groundwater from Ireland. Science of The Total Environment, 705, 135834. https://doi.org/10.1016/j.scitotenv.2019.135834 He, X. S., Xi, B. D., Wei, Z. M., Jiang, Y. H., Geng, C. M., Yang, Y., & Liu, H. L. (2011). Physicochemical and spectroscopic characteristics of dissolved organic matter extracted from municipal solid waste (MSW) and their influence on the landfill biological stability. Bioresource Technology, 102(3), 2322-2327. http://dx.doi.org/10.1016/j.biortech.2010.10.085 He, Y. T., Wilson, J. T., Su, C., & Wilkin, R. T. (2015). Review of abiotic degradation of chlorinated solvents by reactive iron minerals in aquifers. Groundwater Monitoring & Remediation, 35(3), 57-75. https://doi.org/10.1111/gwmr.12111 Helbling, D. E., Johnson, D. R., Honti, M., & Fenner, K. (2012). Micropollutant biotransformation kinetics associate with WWTP process parameters and microbial community characteristics. Environmental Science & Technology, 46(19), 10579-10588. https://doi.org/10.1021/es3019012 Helms, J. R., Stubbins, A., Ritchie, J. D., Minor, E. C., Kieber, D. J., & Mopper, K. (2008). Absorption spectral slopes and slope ratios as indicators of molecular weight, source, and photobleaching of chromophoric dissolved organic matter. Limnology and Oceanography, 53(3), 955-969. https://doi.org/10.4319/lo.2008.53.3.0955 Hepburn, E., Madden, C., Szabo, D., Coggan, T. L., Clarke, B., & Currell, M. (2019). Contamination of groundwater with per-and polyfluoroalkyl substances (PFAS) from legacy landfills in an urban re-development precinct. Environmental Pollution, 248, 101-113. https://doi.org/10.1016/j.envpol.2019.02.018 122  Herzke, D., Olsson, E., Posner, S., (2012). Perfluoroalkyl and polyfluoroalkyl substances (PFASs) in consumer products in Norway–a pilot study. Chemosphere. 88(8), 980-987. https://doi.org/10.1016/j.chemosphere.2012.03.035 Higgins, C. P., & Luthy, R. G. (2006). Sorption of perfluorinated surfactants on sediments. Environmental Science & Technology, 40(23), 7251-7256. https://doi.org/10.1021/es061000n Hill, T. C., Walsh, K. A., Harris, J. A., & Moffett, B. F. (2003). Using ecological diversity measures with bacterial communities. FEMS Microbiology Ecology, 43(1), 1-11. https://doi.org/10.1111/j.1574-6941.2003.tb01040.x Houde, M., De Silva, A. O., Muir, D. C., & Letcher, R. J. (2011). Monitoring of perfluorinated compounds in aquatic biota: an updated review: PFCs in aquatic biota. Environmental Science & Technology, 45(19), 7962-7973. https://doi.org/dx.doi.org/10.1021/es104326w Huang, M., Hu, H., Ma, L., Zhou, Q., Yu, L., & Zeng, S. (2014). Carbon–carbon double-bond reductases in nature. Drug Metabolism Reviews, 46(3), 362-378. https://doi.org/10.3109/03602532.2014.910219 Huo, S., Xi, B., Yu, H., He, L., Fan, S., & Liu, H. (2008). Characteristics of dissolved organic matter (DOM) in leachate with different landfill ages. Journal of Environmental Sciences, 20(4), 492-498.  Huset, C.A., Barlaz, M.A., Barofsky, D.F., Field, J.A., 2011. Quantitative determination of fluorochemicals in municipal landfill leachates. Chemosphere. 82, 1380-1386. https://doi.org/10.1016/j.chemosphere.2010.11.072 Imai, R., Nagata, Y., Senoo, K., Wada, H., Fukuda, M., Takagi, M., & Yano, K. (1989). Dehydrochlorination of γ-hexachlorocyclohexane (γ-BHC) by γ-BHC-assimilating Pseudomonas paucimobilis. Agricultural and biological chemistry, 53(7), 2015-2017. https://doi.org/10.1080/00021369.1989.10869597 Ishibashi, H., Yamauchi, R., Matsuoka, M., Kim, J. W., Hirano, M., Yamaguchi, A., & Arizono, K. (2008). Fluorotelomer alcohols induce hepatic vitellogenin through activation of the estrogen receptor in male medaka (Oryzias latipes). Chemosphere, 71(10), 1853-1859. https://doi.org/10.1016/j.chemosphere.2008.01.065 Ismail, S. N. S., & Latifah, A. M. (2013). The challenge of future landfill: A case study of Malaysia. Journal Toxicology and Environmental Health Sciences, 5(3), 2400-2407. https://doi.org/10.5897/JTEHS12.20  Jahnke, A., Ahrens, L., Ebinghaus, R., & Temme, C. (2007). Urban versus remote air concentrations of fluorotelomer alcohols and other polyfluorinated alkyl substances in Germany. Environmental Science & Technology, 41(3), 745-752. https://doi.org/10.1021/es0619861 Jiang, H., Chen, Y., Jiang, P., Zhang, C., Smith, T. J., Murrell, J. C., & Xing, X. H. (2010). Methanotrophs: multifunctional bacteria with promising applications in environmental bioengineering. Biochemical Engineering Journal, 49(3), 277-288. https://doi.org/10.1016/j.bej.2010.01.003 123  Jiang, Y., Li, Q., Chen, X., & Jiang, C. (2016). Isolation and cultivation methods of Actinobacteria. In Actinobacteria-Basics and Biotechnological Applications. IntechOpen. http://dx.doi.org/10.5772/61457 Johnson, E. (2019). Per-and Polyfluoroalkyl Substances (PFASs) Removal from Landfill Leachate: Efficiency Evaluation in Column Experiments. Degree Project in Sustainable Technology, Environment Department of Sustainable Development, Sweden Juhasz, A. L., & Naidu, R. (2000). Bioremediation of high molecular weight polycyclic aromatic hydrocarbons: a review of the microbial degradation of benzo [a] pyrene. International Biodeterioration & Biodegradation, 45(1-2), 57-88. https://doi.org/10.1016/S0964-8305(00)00052-4 Kaestner, M., Nowak, K. M., Miltner, A., Trapp, S., & Schaeffer, A. (2014). Classification and modelling of nonextractable residue (NER) formation of xenobiotics in soil–a synthesis. Critical Reviews in Environmental Science and Technology, 44(19), 2107-2171. https://doi.org/10.1080/10643389.2013.828270 Kallenborn, R.K., Berger, U., Jarnberg, U., (2004). Perfluorinated Alkylated Substances (PFAS) in the Nordic Environment. Nordic Council of Ministers. Kannan, K., Corsolini, S., Falandysz, J., Fillmann, G., Kumar, K. S., Loganathan, B. G., & Aldous, K. M. (2004). Perfluorooctanesulfonate and related fluorochemicals in human blood from several countries. Environmental Science & Technology, 38(17), 4489-4495. https://doi.org/10.1021/es0493446 Karigar, C. S., & Rao, S. S. (2011). Role of microbial enzymes in the bioremediation of pollutants: a review. Enzyme Research, 2011. http://dx.doi.org/10.4061/2011/805187 Kästner, M., & Richnow, H. H. (2001). Formation of residues of organic pollutants within the soil matrix—Mechanisms and stability. In Treatment of contaminated soil (pp. 219-251). Springer, Berlin, Heidelberg. https://doi.org/10.1007/978-3-662-04643-2_16 Kato, K., Wong, L. Y., Jia, L. T., Kuklenyik, Z., & Calafat, A. M. (2011). Trends in exposure to polyfluoroalkyl chemicals in the US population: 1999− 2008. Environmental Science & Technology, 45(19), 8037-8045. https://doi.org/10.1021/es1043613 Kaza, S., Yao, L., Bhada-Tata, P., & Van Woerden, F. (2018). What a waste 2.0: a global snapshot of solid waste management to 2050. World Bank Publications. Available at http://datatopics.worldbank.org/what-a-waste/trends_in_solid_waste_management.html. Accessed on October, 2019 Keener, W. K., & Arp, D. J. (1993). Kinetic studies of ammonia monooxygenase inhibition in Nitrosomonas europaea by hydrocarbons and halogenated hydrocarbons in an optimized whole-cell assay. Applied and Environmental Microbiology, 59(8), 2501-2510. Key, B. D., Howell, R. D., & Criddle, C. S. (1998). Defluorination of organofluorine sulfur compounds by Pseudomonas sp. strain D2. Environmental Science & Technology, 32(15), 2283-2287. http://dx.doi.org/10.1021/es9800129 Khosravifarsani, M., Shabestani-Monfared, A., Pouramir, M., & Zabihi, E., (2016). Hydroxyl Radical (ºOH) Scavenger Power of Tris (hydroxymethyl) Compared to Phosphate Buffer. Journal of Molecular Biology Research, 6 (1).  124  Khunjar, W. O., Mackintosh, S. A., Skotnicka-Pitak, J., Baik, S., Aga, D. S., & Love, N. G. (2011). Elucidating the relative roles of ammonia oxidizing and heterotrophic bacteria during the biotransformation of 17α-ethinylestradiol and trimethoprim. Environmental Science & Technology, 45(8), 3605-3612. http://dx.doi.org/10.1021/es1037035 Kielak, A. M., Barreto, C. C., Kowalchuk, G. A., van Veen, J. A., & Kuramae, E. E. (2016). The ecology of Acidobacteria: moving beyond genes and genomes. Frontiers in Microbiology, 7, 744. https://doi.org/10.3389/fmicb.2016.00744 Kim, M. H., Wang, N., & Chu, K. H. (2014). 6: 2 Fluorotelomer alcohol (6: 2 FTOH) biotransformation by multiple microbial species under different physiological conditions. Applied microbiology and Biotechnology, 98(4), 1831-1840. https://doi.org/10.1007/s00253-013-5131-3 Kim, M. H., Wang, N., McDonald, T., & Chu, K. H. (2012). Biodefluorination and biotransformation of fluorotelomer alcohols by two alkane‐degrading Pseudomonas strains. Biotechnology and Bioengineering, 109(12), 3041-3048. https://doi.org/10.1002/bit.24561 Kim, M., Li, L. Y., Grace, J. R., Benskin, J. P., & Ikonomou, M. G. (2015). Compositional effects on leaching of stain-guarded (perfluoroalkyl and polyfluoroalkyl substance-treated) carpet in landfill leachate. Environmental Science & Technology, 49(11), 6564-6573. https://doi.org/10.1021/es505333y Kissa, E. (Ed.). (2001). Fluorinated surfactants and repellents (Vol. 97). CRC Press. Kjeldsen, P., Barlaz, M. A., Rooker, A. P., Baun, A., Ledin, A., & Christensen, T. H. (2002). Present and long-term composition of MSW landfill leachate: a review. Critical Reviews in Environmental Science and Technology, 32(4), 297-336. https://doi.org/10.1080/10643380290813462 Knapp, J. S., & Bromley-Challoner, K. C. (2003). Recalcitrant organic compounds. Handbook of Water and Wastewater Microbiology. Academic Press, London, 559-595. https://doi.org/10.1016/B978-012470100-7/50035-2 Knief, C. (2015). Diversity and habitat preferences of cultivated and uncultivated aerobic methanotrophic bacteria evaluated based on pmoA as molecular marker. Frontiers in Microbiology, 6, 1346. https://doi.org/10.3389/fmicb.2015.01346 Knutsen, H., Mæhlum, T., Haarstad, K., Slinde, G. A., & Arp, H. P. H. (2019). Leachate emissions of short-and long-chain per-and polyfluoralkyl substances (PFASs) from various Norwegian landfills. Environmental Science: Processes & Impacts, 21(11), 1970-1979. https://doi.org/10.1039/c9em00170k Köchling, T., Sanz, J. L., Gavazza, S., & Florencio, L. (2015). Analysis of microbial community structure and composition in leachates from a young landfill by 454 pyrosequencing. Applied Microbiology and Biotechnology, 99(13), 5657-5668. https://doi.org/10.1007/s00253-015-6409-4 Kong, B., Wang, X., He, B., Wei, L., Zhu, J., Jin, Y., & Fu, Z. (2019). 8: 2 fluorotelomer alcohol inhibited proliferation and disturbed the expression of pro-inflammatory cytokines 125  and antigen-presenting genes in murine macrophages. Chemosphere, 219, 1052-1060. https://doi.org/10.1016/j.chemosphere.2018.12.091 Kotthoff, M., Müller, J., Jürling, H., Schlummer, M., & Fiedler, D. (2015). Perfluoroalkyl and polyfluoroalkyl substances in consumer products. Environmental Science and Pollution Research, 22(19), 14546-14559. https://doi.org/10.1007/s11356-015-4202-7. Kozich, J. J., Westcott, S. L., Baxter, N. T., Highlander, S. K., & Schloss, P. D. (2013). Development of a dual-index sequencing strategy and curation pipeline for analyzing amplicon sequence data on the MiSeq Illumina sequencing platform. Applied and Environmental Microbiology, AEM-01043. https://doi.org/10.1128/AEM.01043-13 Krafft, M. P., & Riess, J. G. (2015). Selected physicochemical aspects of poly-and perfluoroalkylated substances relevant to performance, environment and sustainability—Part one. Chemosphere, 129, 4-19. https://doi.org/10.1016/j.chemosphere.2014.08.039 Krivoruchko, A., Kuyukina, M., & Ivshina, I. (2019). Advanced Rhodococcus biocatalysts for environmental biotechnologies. Catalysts, 9(3), 236. https://doi.org/10.3390/catal9030236 Kulikowska, D., & Klimiuk, E. (2008). The effect of landfill age on municipal leachate composition. Bioresource Technology, 99(13), 5981-5985. https://doi.org/10.1016/j.biortech.2007.10.015 Kuppusamy, S., Thavamani, P., Megharaj, M., Lee, Y. B., & Naidu, R. (2016). Polyaromatic hydrocarbon (PAH) degradation potential of a new acid tolerant, diazotrophic P-solubilizing and heavy metal resistant bacterium Cupriavidus sp. MTS-7 isolated from long-term mixed contaminated soil. Chemosphere, 162, 31-39. https://doi.org/10.1016/j.chemosphere.2016.07.052 Kwok, K. Y., Wang, X. H., Ya, M., Li, Y., Zhang, X. H., Yamashita, N., & Lam, P. K. (2015). Occurrence and distribution of conventional and new classes of per-and polyfluoroalkyl substances (PFASs) in the South China Sea. Journal of Hazardous Materials, 285, 389-397. https://doi.org/10.1016/j.jhazmat.2014.10.065 Lang, J. R., Allred, B. M., Field, J. A., Levis, J. W., & Barlaz, M. A. (2017). National estimate of per-and polyfluoroalkyl substance (PFAS) release to US municipal landfill leachate. Environmental Science & Technology, 51(4), 2197-2205. https://doi.org/10.1021/acs.est.6b05005 Lang, J. R., Allred, B. M., Peaslee, G. F., Field, J. A., & Barlaz, M. A. (2016). Release of per-and polyfluoroalkyl substances (PFASs) from carpet and clothing in model anaerobic landfill reactors. Environmental Science & Technology, 50(10), 5024-5032. https://doi.org/10.1021/acs.est.5b06237 Lassen, C., Jensen, A. A., Potrykus, A., Christensen, F., Kjolholt, J., Jeppesen, C. N., & Innanen, S. (2013). Survey of PFOS, PFOA and other perfluoroalkyl and polyfluoroalkyl substances. Part of the LOUS-review Environmental Project, (1475). Lee, H., D’eon, J., & Mabury, S. A. (2010). Biodegradation of polyfluoroalkyl phosphates as a source of perfluorinated acids to the environment. Environmental Science & Technology, 44(9), 3305-3310. https://doi.org/10.1021/es9028183 126  Lee, H., Tevlin, A. G., Mabury, S. A., & Mabury, S. A. (2014). Fate of polyfluoroalkyl phosphate diesters and their metabolites in biosolids-applied soil: biodegradation and plant uptake in greenhouse and field experiments. Environmental Science & Technology, 48(1), 340-349. https://doi.org/10.1021/es403949z Lewin, G. R., Carlos, C., Chevrette, M. G., Horn, H. A., McDonald, B. R., Stankey, R. J., & Currie, C. R. (2016). Evolution and ecology of Actinobacteria and their bioenergy applications. Annual Review of Microbiology, 70, 235-254. https://doi.org/10.1146/annurev-micro-102215-095748 Lewis, M., Kim, M. H., Wang, N., & Chu, K. H. (2016). Engineering artificial communities for enhanced FTOH degradation. Science of The Total Environment, 572, 935-942. https://doi.org/10.1016/j.scitotenv.2016.07.223 Li, T., Mazéas, L., Sghir, A., Leblon, G., & Bouchez, T. (2009). Insights into networks of functional microbes catalysing methanization of cellulose under mesophilic conditions. Environmental Microbiology, 11(4), 889-904. https://doi.org/10.1111/j.1462-2920.2008.01810.x Li, B. (2011). Perfluorinated compounds in landfill leachate and their effect on the performance of sodium bentonite landfill liners. Masters Dissertation. The University of British Columbia, Canada Li, B., Danon-Schaffer, M. N., Li, L. Y., Ikonomou, M. G., & Grace, J. R. (2012). Occurrence of PFCs and PBDEs in landfill leachates from across Canada. Water, Air, & Soil Pollution, 223(6), 3365-3372. https://doi.org/10.1007/s11270-012-1115-7 Li, L., Zhai, Z., Liu, J., Hu, J. (2015a). Estimating industrial and domestic environmental releases of perfluorooctanoic acid and its salts in China from 2004 to 2012. Chemosphere, 129, 100-109. http://dx.doi.org/10.1016/j.chemosphere.2014.11.049. Li, B., Li, L. Y., & Grace, J. R. (2015b). Adsorption and hydraulic conductivity of landfill-leachate perfluorinated compounds in bentonite barrier mixtures. Journal of Environmental Management, 156, 236-243. https://doi.org/10.1016/j.jenvman.2015.04.003 Li, Z. M., Guo, L. H., & Ren, X. M. (2016). Biotransformation of 8: 2 fluorotelomer alcohol by recombinant human cytochrome P450s, human liver microsomes and human liver cytosol. Environmental Science: Processes & Impacts, 18(5), 538-546. https://doi.org/10.1039/C6EM00071A Li, L., Liu, J., Hu, J., & Wania, F. (2017). Degradation of fluorotelomer-based polymers contributes to the global occurrence of fluorotelomer alcohol and perfluoroalkyl carboxylates: A combined dynamic substance flow and environmental fate modeling analysis. Environmental Science & Technology, 51(8), 4461-4470. https://doi.org/10.1021/acs.est.6b04021 Li, F., Duan, J., Tian, S., Ji, H., Zhu, Y., Wei, Z., & Zhao, D. (2019). Short-chain Per-and Polyfluoroalkyl Substances in Aquatic Systems: Occurrence, Impacts and Treatment. Chemical Engineering Journal, 122506. https://doi.org/10.1016/j.cej.2019.122506 127  Lindstrom, A.B., Strynar, M.J., Libelo, E.L., (2011). Polyfluorinated compounds: past, present, and future. Environmental Science & Technology, 45, 7954-7961. https://doi.org/10.1021/es2011622 Liu, J., & Mejia-Avendaño, S. (2013). Microbial degradation of polyfluoroalkyl chemicals in the environment: a review. Environment International, 61, 98-114. https://doi.org/10.1016/j.envint.2013.08.022 Liu, J., & Lee, L. S. (2005). Solubility and sorption by soils of 8: 2 fluorotelomer alcohol in water and cosolvent systems. Environmental Science & Technology, 39(19), 7535-7540. https://doi.org/10.1021/es051125c Liu, C., & Liu, J. (2016). Aerobic biotransformation of polyfluoroalkyl phosphate esters (PAPs) in soil. Environmental Pollution, 212, 230-237. https://doi.org/10.1016/j.envpol.2016.01.069 Liu, J., Lee, L. S., Nies, L. F., Nakatsu, C. H., & Turco, R. F. (2007). Biotransformation of 8: 2 fluorotelomer alcohol in soil and by soil bacteria isolates. Environmental Science & Technology, 41(23), 8024-8030. https://doi.org/10.1021/es0708722 Liu, C., Yu, L., Deng, J., Lam, P. K., Wu, R. S., & Zhou, B. (2009). Waterborne exposure to fluorotelomer alcohol 6: 2 FTOH alters plasma sex hormone and gene transcription in the hypothalamic–pituitary–gonadal (HPG) axis of zebrafish. Aquatic Toxicology, 93(2-3), 131-137. https://doi.org/10.1016/j.aquatox.2009.04.005 Liu, C., Deng, J., Yu, L., Ramesh, M., & Zhou, B. (2010a). Endocrine disruption and reproductive impairment in zebrafish by exposure to 8: 2 fluorotelomer alcohol. Aquatic Toxicology, 96(1), 70-76. https://doi.org/10.1016/j.aquatox.2009.09.012 Liu, C., Zhang, X., Chang, H., Jones, P., Wiseman, S., Naile, J., & Zhou, B. (2010b). Effects of fluorotelomer alcohol 8: 2 FTOH on steroidogenesis in H295R cells: targeting the cAMP signalling cascade. Toxicology and Applied Pharmacology, 247(3), 222-228. https://doi.org/10.1016/j.taap.2010.06.016 Liu, J., Wang, N., Szostek, B., Buck, R. C., Panciroli, P. K., Folsom, P. W., & Bellin, C. A. (2010c). 6-2 Fluorotelomer alcohol aerobic biotransformationin soil and mixed bacterial culture. Chemosphere, 78(4), 437-444. https://doi.org/10.1016/j.chemosphere.2013.02.032 Liu, X., Guo, Z., Krebs, K. A., Pope, R. H., & Roache, N. F. (2014). Concentrations and trends of perfluorinated chemicals in potential indoor sources from 2007 through 2011 in the US. Chemosphere, 98, 51-57. https://doi.org/10.1016/j.chemosphere.2013.10.001 Liu, X., Guo, Z., Folk IV, E. E., & Roache, N. F. (2015). Determination of fluorotelomer alcohols in selected consumer products and preliminary investigation of their fate in the indoor environment. Chemosphere, 129, 81-86. http://dx.doi.org/10.1016/j.chemosphere.2014.06.012 Liu, X., Shu, Z., Sun, D., Dang, Y., & Holmes, D. E. (2018). Heterotrophic nitrifiers dominate reactors treating incineration leachate with high free ammonia concentrations. ACS Sustainable Chemistry & Engineering, 6(11), 15040-15049. http://dx.doi.org/10.1021/acssuschemeng.8b03512 128  Maizel, A. C. (2017). Relating Dissolved Organic Matter Composition and Photochemistry with High Resolution Mass Spectrometry. Doctoral dissertation, The University of Wisconsin-Madison. Maizel, A. C., & Remucal, C. K. (2017). The effect of advanced secondary municipal wastewater treatment on the molecular composition of dissolved organic matter. Water Research, 122, 42-52. https://doi.org/10.1016/j.watres.2017.05.055 Maras, M., Vanparys, C., Muylle, F., Robbens, J., Berger, U., Barber, J. L., & De Coen, W. (2006). Estrogen-like properties of fluorotelomer alcohols as revealed by MCF-7 breast cancer cell proliferation. Environmental Health Perspectives, 114(1), 100. https://doi.org/10.1289/ehp.8149  Marchant, H. K., Ahmerkamp, S., Lavik, G., Tegetmeyer, H. E., Graf, J., Klatt, J. M., & Kuypers, M. M. (2017). Denitrifying community in coastal sediments performs aerobic and anaerobic respiration simultaneously. The ISME Journal, 11(8), 1799. https://doi.org/10.1038/ismej.2017.51 McKay, G., Couch, K. D., Mezyk, S. P., & Rosario-Ortiz, F. L. (2016). Investigation of the coupled effects of molecular weight and charge-transfer interactions on the optical and photochemical properties of dissolved organic matter. Environmental Science & Technology, 50(15), 8093-8102. https://doi.org/10.1021/acs.est.6b02109 Mejia-Avendaño, S., Vo Duy, S., Sauvé, S., & Liu, J. (2016). Generation of perfluoroalkyl acids from aerobic biotransformation of quaternary ammonium polyfluoroalkyl surfactants. Environmental Science & Technology, 50(18), 9923-9932. https://doi.org/10.1021/acs.est.6b00140 Men, Y., Achermann, S., Helbling, D. E., Johnson, D. R., & Fenner, K. (2017). Relative contribution of ammonia oxidizing bacteria and other members of nitrifying activated sludge communities to micropollutant biotransformation. Water Research, 109, 217-226. https://doi.org/10.1016/j.watres.2016.11.048 Merino, N., Qu, Y., Deeb, R. A., Hawley, E. L., Hoffmann, M. R., & Mahendra, S. (2016). Degradation and removal methods for perfluoroalkyl and polyfluoroalkyl substances in water. Environmental Engineering Science, 33(9), 615-649. https://doi.org/10.1089/ees.2016.0233 Mitchell, R. J., Myers, A. L., Mabury, S. A., Solomon, K. R., & Sibley, P. K. (2011). Toxicity of fluorotelomer carboxylic acids to the algae Pseudokirchneriella subcapitata and Chlorella vulgaris, and the amphipod Hyalella azteca. Ecotoxicology and Environmental Safety, 74(8), 2260-2267. https://doi.org/10.1016/j.ecoenv.2011.07.034 Moir, J. W., Crossman, L. C., Spiro, S., & Richardson, D. J. (1996). The purification of ammonia monooxygenase from Paracoccus denitrficans. FEBS letters, 387(1), 71-74. https://doi.org/10.1016/0014-5793(96)00463-2 Mwaikono, K. S., Maina, S., Sebastian, A., Schilling, M., Kapur, V., & Gwakisa, P. (2016). High-throughput sequencing of 16S rRNA gene reveals substantial bacterial diversity on the municipal dumpsite. BMC Microbiology, 16(1), 145. https://doi.org/10.1186/s12866-016-0758-8 129  Mylchreest, E., Munley, S. M., & Kennedy Jr, G. L. (2005). Evaluation of the developmental toxicity of 8-2 telomer B alcohol. Drug and chemical toxicology, 28(3), 315-328. https://doi.org/10.1081/DCT-200064491 Naether, A., Foesel, B. U., Naegele, V., Wüst, P. K., Weinert, J., Bonkowski, M., & Gockel, S. (2012). Environmental factors affect acidobacterial communities below the subgroup level in grassland and forest soils. Applied Environmental Microbiology., 78(20), 7398-7406. https://doi.org/10.1128/AEM.01325-12 Nagata, Y., Hatta, T., Imai, R., Kimbara, K., Fukuda, M., Yano, K., & Takagi, M. (1993). Purification and characterization of γ-hexachlorocyclohexane (γ-HCH) dehydrochlorinase (LinA) from Pseudomonas paucimobilis. Bioscience, Biotechnology, and Biochemistry, 57(9), 1582-1583. https://doi.org/10.1271/bbb.57.1582 Newton, R. J., Jones, S. E., Eiler, A., McMahon, K. D., & Bertilsson, S. (2011). A guide to the natural history of freshwater lake bacteria. Microbiology and Molecular Biology Reviews, 75(1), 14-49. https://doi.org/10.1128/MMBR.00028-10 Nogueira, R., & Melo, L. F. (2006). Competition between Nitrospira spp. and Nitrobacter spp. in nitrite‐oxidizing bioreactors. Biotechnology and Bioengineering, 95(1), 169-175.  https://doi.org/10.1002/bit.21004 Nowka, B., Daims, H., & Spieck, E. (2015). Comparison of oxidation kinetics of nitrite-oxidizing bacteria: nitrite availability as a key factor in niche differentiation. Applied and Environmental Microbiology, 81(2), 745-753. https://doi.org/10.1128/AEM.02734-14 Nsenga Kumwimba, M., & Meng, F. (2019). Roles of ammonia-oxidizing bacteria in improving metabolism and cometabolism of trace organic chemicals in biological wastewater treatment processes: A review. Science of the Total Environment. https://doi.org/10.1016/j.scitotenv.2018.12.236 Ochoa-Herrera, V., Field, J. A., Luna-Velasco, A., & Sierra-Alvarez, R. (2016). Microbial toxicity and biodegradability of perfluorooctane sulfonate (PFOS) and shorter chain perfluoroalkyl and polyfluoroalkyl substances (PFASs). Environmental Science: Processes & Impacts, 18(9), 1236-1246. https://doi.org/10.1039/C6EM00366D OECD, Organization for Economic Cooperation and Development, (1992). OECD Guideline For Testing Of Chemicals. Available from: https://www.oecd.org/env/ehs/testing/oecdguidelinesforthetestingofchemicals.htm Accessed on December, 2019 Oliaei, F., Kriens, D., Kessler, K., (2006). Investigation of Perfluorochemical (PFC) Contamination in Minnesota, Phase One: Report to Senate Environment Committee. Oturan, N., Van Hullebusch, E. D., Zhang, H., Mazeas, L., Budzinski, H., Le Menach, K., & Oturan, M. A. (2015). Occurrence and removal of organic micropollutants in landfill leachates treated by electrochemical advanced oxidation processes. Environmental Science & Technology, 49(20), 12187-12196. https://doi.org/10.1021/acs.est.5b02809 Pan, C. G., Liu, Y. S., & Ying, G. G. (2016). Perfluoroalkyl substances (PFASs) in wastewater treatment plants and drinking water treatment plants: Removal efficiency and exposure risk. Water Research, 106, 562-570. https://doi.org/10.1016/j.watres.2016.10.045 130  Parales, R. E. (2010). Hydrocarbon degradation by Betaproteobacteria. Handbook of Hydrocarbon and Lipid Microbiology, 1715-1724. https://doi.org/10.1007/978-3-540-77587-4_121 Parker, K. M., Pignatello, J. J., & Mitch, W. A. (2013). Influence of ionic strength on triplet-state natural organic matter loss by energy transfer and electron transfer pathways. Environmental Science & Technology, 47(19), 10987-10994. https://doi.org/10.1021/es401900j Paul, D., Pandey, G., Meier, C., Roelof van der Meer, J., & Jain, R. K. (2006). Bacterial community structure of a pesticide-contaminated site and assessment of changes induced in community structure during bioremediation. FEMS Microbiology Ecology, 57(1), 116-127. https://doi.org/10.1111/j.1574-6941.2006.00103.x Perkola, N., & Sainio, P. (2013). Survey of perfluorinated alkyl acids in Finnish effluents, storm water, landfill leachate and sludge. Environmental Science and Pollution Research, 20(11), 7979-7987. https://doi.org/10.1007/s11356-013-1518-z Phillips, M. M., Dinglasan-Panlilio, M. J. A., Mabury, S. A., Solomon, K. R., & Sibley, P. K. (2007). Fluorotelomer acids are more toxic than perfluorinated acids. Environmental science & technology, 41(20), 7159-7163. https://doi.org/10.1021/es070734c Phillips, M. M., Dinglasan‐Panlilio, M. J., Mabury, S. A., Solomon, K. R., & Sibley, P. K. (2010). Chronic toxicity of fluorotelomer acids to Daphnia magna and Chironomus dilutus. Environmental Toxicology and Chemistry, 29(5), 1123-1131. https://doi.org/10.1002/etc.141 Plósz, B. G., Leknes, H., & Thomas, K. V. (2009). Impacts of competitive inhibition, parent compound formation and partitioning behavior on the removal of antibiotics in municipal wastewater treatment. Environmental Science & Technology, 44(2), 734-742. https://doi.org/10.1021/es902264w Poulsen, P. B., Gram, L. K., Jensen, A. A., Rasmussen, A. A., Ravn, C., Møller, P., & Løkkegaard, K. (2011). Substitution of PFOS for use in non-decorative hard chrome plating. Washington: Environmental Protection Agency. Qiao, W., Xie, Z., Zhang, Y., Liu, X., Xie, S., Huang, J., & Yu, L. (2018). Perfluoroalkyl substances (PFASs) influence the structure and function of soil bacterial community: Greenhouse experiment. Science of The Total Environment, 642, 1118-1126. https://doi.org/10.1016/j.scitotenv.2018.06.113 Rahman, M. F., Peldszus, S., & Anderson, W. B. (2014). Behaviour and fate of perfluoroalkyl and polyfluoroalkyl substances (PFASs) in drinking water treatment: a review. Water Research, 50, 318-340. https://doi.org/10.1016/j.watres.2013.10.045 Rand, A. A., Rooney, J. P., Butt, C. M., Meyer, J. N., & Mabury, S. A. (2013). Cellular toxicity associated with exposure to perfluorinated carboxylates (PFCAs) and their metabolic precursors. Chemical Research in Toxicology, 27(1), 42-50 https://doi.org/10.1021/tx400317p Rankin, K., Lee, H., Tseng, P. J., & Mabury, S. A. (2014). Investigating the biodegradability of a fluorotelomer-based acrylate polymer in a soil–plant microcosm by indirect and direct 131  analysis. Environmental Science & Technology, 48(21), 12783-12790. https://doi.org/10.1021/es502986w  Rao, N. S., Baker, B. E. (1994). Textile Finishes & Fluorosurfactants. In Organofluorine Chemistry. Principles and Commercial Applications; Banks, R. E., Smart, B. E., Tatlow, J. C., Eds.; Plenum Press: New York; pp 321-336 Rayne, S., & Forest, K. (2009). Perfluoroalkyl sulfonic and carboxylic acids: a critical review of physicochemical properties, levels and patterns in waters and wastewaters, and treatment methods. Journal of Environmental Science and Health Part A, 44(12), 1145-1199. https://doi.org/10.1080/10934520903139811 Remucal, C. K. (2014). The role of indirect photochemical degradation in the environmental fate of pesticides: a review. Environmental Science: Processes & Impacts, 16(4), 628-653. 10.1039/c3em00549f Ren, D., Huang, B., Xiong, D., He, H., Meng, X., & Pan, X. (2017). Photodegradation of 17α-ethynylestradiol in dissolved humic substances solution: Kinetics, mechanism and estrogenicity variation. Journal of Environmental Sciences, 54, 196-205. http://dx.doi.org/10.1016/j.jes.2016.03.002 Ren, T. T., Jin, C. Z., Jin, F. J., Li, T., Kim, C. J., Oh, H. M., & Jin, L. (2018). Flavihumibacter profundi sp. nov., isolated from eutrophic freshwater sediment. Journal of Microbiology, 56(7), 467-471. https://doi.org/10.1007/s12275-018-7567-8 Ren, Y., Ferraz, F., Lashkarizadeh, M., & Yuan, Q. (2017). Comparing young landfill leachate treatment efficiency and process stability using aerobic granular sludge and suspended growth activated sludge. Journal of Water Process Egineering, 17, 161-167. https://doi.org/10.1016/j.jwpe.2017.04.006 Renou, S., Givaudan, J. G., Poulain, S., Dirassouyan, F., & Moulin, P. (2008). Landfill leachate treatment: review and opportunity. Journal of Hazardous Materials, 150(3), 468-493. https://doi.org/10.1016/j.jhazmat.2007.09.077 Rhoads, K. R., Janssen, E. M. L., Luthy, R. G., & Criddle, C. S. (2008). Aerobic biotransformation and fate of N-ethyl perfluorooctane sulfonamidoethanol (N-EtFOSE) in activated sludge. Environmental Science & Technology, 42(8), 2873-2878. https://doi.org/10.1021/es702866c Ritter S. 2010. Fluorochemicals go short. Chemical & Engineering News 88: 12-17. https://doi.org/10.1021/cenv088n005. Robinson, H. (2007). The composition of leachates from very large landfills: an international review. 10th International Waste Management and Landfill Symposium, 8(1), 19-32. S. Cagliari, Italy; 3 - 7 October. Roh, H., Subramanya, N., Zhao, F., Yu, C. P., Sandt, J., & Chu, K. H. (2009). Biotransformation potential of wastewater micropollutants by ammonia-oxidizing bacteria. Chemosphere, 77(8), 1084-1089. https://doi.org/10.1016/j.chemosphere.2009.08.049 Rosenmai, A. K., Nielsen, F. K., Pedersen, M., Hadrup, N., Trier, X., Christensen, J. H., & Vinggaard, A. M. (2013). Fluorochemicals used in food packaging inhibit male sex hormone 132  synthesis. Toxicology and Applied Pharmacology, 266(1), 132-142. https://doi.org/10.1016/j.taap.2012.10.022 Sánchez-peinado, M., González-lópez, J., Martínez-toledo, M. V., Pozo, C., & Rodelas, B. (2010). Influence of linear alkylbenzene sulfonate (LAS) on the structure of Alphaproteobacteria, Actinobacteria, and Acidobacteria communities in a soil microcosm. Environmental Science and Pollution Research International, 17(3), 779. https://doi.org/10.1007/s11356-009-0180-y Sathyamoorthy, S., Chandran, K., & Ramsburg, C. A. (2013). Biodegradation and cometabolic modeling of selected beta blockers during ammonia oxidation. Environmental Science & Technology, 47(22), 12835-12843. https://doi.org/10.1021/es402878e Schloss, P. D., Westcott, S. L., Ryabin, T., Hall, J. R., Hartmann, M., Hollister, E. B., & Sahl, J. W. (2009). Introducing mothur: open-source, platform-independent, community-supported software for describing and comparing microbial communities. Applied and Environmental Microbiology, 75(23), 7537-7541. https://doi.org/10.1128/AEM.01541-09 Schmalenberger, A., Hodge, S., Hawkesford, M. J., & Kertesz, M. A. (2009). Sulfonate desulfurization in Rhodococcus from wheat rhizosphere communities. FEMS Microbiology ecology, 67(1), 140-150. https://doi.org/1111/j.1574-6941.2008.00602.x Scholes, R. C., Prasse, C., & Sedlak, D. L. (2019). The role of reactive nitrogen species in sensitized photolysis of wastewater-derived trace organic contaminants. Environmental Science & Technology, 53(11), 6483-6491. https://10.1021/acs.est.9b01386 Schrijver, A. D., & Mot, R. D. (1999). Degradation of pesticides by actinomycetes. Critical Reviews in Microbiology, 25(2), 85-119. https://doi.org/10.1080/10408419991299194 Schwarzenbach R.P., Gschwend P.M., Imboden D.M., (2003). Environmental Organic  Chemistry. John Wiley & Sons, Inc., Hoboken, NJ. Semrau, J. D. (2011). Current knowledge of microbial community structures in landfills and its cover soils. Applied Microbiology and Biotechnology, 89(4), 961-969. https://doi.org/10.1007/s00253-010-3024-2 Sharpless, C. M. (2012). Lifetimes of triplet dissolved natural organic matter (DOM) and the effect of NaBH4 reduction on singlet oxygen quantum yields: Implications for DOM photophysics. Environmental Science & Technology, 46(8), 4466-4473. https://doi.org/10.1021/es300217h Shaw, D. M., Munoz, G., Bottos, E. M., Duy, S. V., Sauvé, S., Liu, J., & Van Hamme, J. D. (2019). Degradation and defluorination of 6: 2 fluorotelomer sulfonamidoalkyl betaine and 6: 2 fluorotelomer sulfonate by Gordonia sp. strain NB4-1Y under sulfur-limiting conditions. Science of The Total Environment, 647, 690-698. https://doi.org/10.1016/j.scitotenv.2018.08.012 Shi, G., Cui, Q., Pan, Y., Sheng, N., Guo, Y., & Dai, J. (2017). 6: 2 fluorotelomer carboxylic acid (6: 2 FTCA) exposure induces developmental toxicity and inhibits the formation of erythrocytes during zebrafish embryogenesis. Aquatic Toxicology, 190, 53-61. https://doi.org/10.1016/j.aquatox.2017.06.023 133  Shoeib, M., Harner, T., & Vlahos, P. (2006). Perfluorinated chemicals in the Arctic atmosphere. Environmental Science & Technology, 40(24), 7577-7583.  https://doi.org/10.1021/es0618999 Song, L., Wang, Y., Zhao, H., & Long, D. T. (2015). Composition of bacterial and archaeal communities during landfill refuse decomposition processes. Microbiological Research, 181, 105-111. https://doi.org/10.1016/j.micres.2015.04.009 Stamps, B. W., Lyles, C. N., Suflita, J. M., Masoner, J. R., Cozzarelli, I. M., Kolpin, D. W., & Stevenson, B. S. (2016). Municipal solid waste landfills harbor distinct microbiomes. Frontiers in Microbiology, 7, 534. https://doi.org/10.3389/fmicb.2016.00534 Stevens, H., Stübner, M., Simon, M., & Brinkhoff, T. (2005). Phylogeny of Proteobacteria and Bacteroidetes from oxic habitats of a tidal flat ecosystem. FEMS Microbiology Ecology, 54(3), 351-365. https://doi.org/10.1016/j.femsec.2005.04.008 Stockholm Convention (2016). Persistent Organic Pollutant Review Committee Twelfth Meeting. Available at http://chm.pops.int/Convention/POPsReviewCommittee/Chemicals/tabid/243/Default.aspx. Accessed on July, 2017. Stockholm Convention (2018). Information on PFOA, its salts and PFOA-related compounds. Available at http://chm.pops.int/TheConvention/POPsReviewCommittee/Meetings/POPRC13/POPRC13Followup/PFOAInfoSubmission/tabid/6174/Default.aspx. Accessed on November 2018 Stockholm Convention (2019). Compilation of comments received from Parties relating to the listing of chemicals in Annexes A, B and/or C to the Stockholm Convention recommended by the Persistent Organic Pollutants Review Committee. Available at. http://www.brsmeas.org/2019COPs/MeetingDocuments/tabid/7832/language/en-US/Default.aspx .Accessed on June, 2019 Sui, Q., Zhao, W., Cao, X., Lu, S., Qiu, Z., Gu, X., & Yu, G. (2017). Pharmaceuticals and personal care products in the leachates from a typical landfill reservoir of municipal solid waste in Shanghai, China: Occurrence and removal by a full-scale membrane bioreactor. Journal of Hazardous Materials, 323, 99-108. http://dx.doi.org/10.1016/j.jhazmat.2016.03.047 Sun, Y., Wang, T., Peng, X., Wang, P., & Lu, Y. (2016). Bacterial community compositions in sediment polluted by perfluoroalkyl acids (PFAAs) using Illumina high-throughput sequencing. Environmental Science and Pollution Research, 23(11), 10556-10565. https://doi.org/10.1007/s11356-016-6055-0 Sutton, N. B., Maphosa, F., Morillo, J. A., Al-Soud, W. A., Langenhoff, A. A., Grotenhuis, T., & Smidt, H. (2013). Impact of long-term diesel contamination on soil microbial community structure. Applied and Environmental Microbiology, 79(2), 619-630. https://doi.org/10.1128/AEM.02747-12 Tobiszewski, M., & Namieśnik, J. (2012). Abiotic degradation of chlorinated ethanes and ethenes in water. Environmental Science and Pollution Research, 19(6), 1994-2006. https://doi.org/10.1007/s11356-012-0764-9 134  Tortoso, A. C., & Hutchinson, G. L. (1990). Contributions of autotrophic and heterotrophic nitrifiers to soil NO and N2O emissions. Applied and Environmental Microbiology, 56(6), 1799-1805. Townsend, T. G., Powell, J., Jain, P., Xu, Q., Tolaymat, T., & Reinhart, D. (2015). Sustainable practices for landfill design and operation. Springer. http://dx.doi.org/10.1007/978-1-4939-2662-6 Tran, N. H., Urase, T., & Kusakabe, O. (2009). The characteristics of enriched nitrifier culture in the degradation of selected pharmaceutically active compounds. Journal of Hazardous Materials, 171(1-3), 1051-1057. https://doi.org/10.1016/j.jhazmat.2009.06.114 Tran, N. H., Urase, T., Ngo, H. H., Hu, J., & Ong, S. L. (2013). Insight into metabolic and cometabolic activities of autotrophic and heterotrophic microorganisms in the biotransformation of emerging trace organic contaminants. Bioresource technology, 146, 721-731. https://doi.org/10.1016/j.biortech.2013.07.083 Trouborst, L. (2016). Aqueous photolysis of 6: 2 fluorotelomer sulfonamide alkylbetaine. Masters Dissertation, University of Toronto. Tsarpali, V., Kamilari, M., & Dailianis, S. (2012). Seasonal alterations of landfill leachate composition and toxic potency in semi-arid regions. Journal of hazardous materials, 233, 163-171. http://dx.doi.org/10.1016/j.jhazmat.2012.07.007 USEPA (2011). Draft Procedure for Analysis of Perfluorinated Carboxylic Acids and Sulfonic Acids in Sewage Sludge and Biosolids by HPLC/MS/MS.  Van Beilen, J. B., & Funhoff, E. G. (2007). Alkane hydroxylases involved in microbial alkane degradation. Applied Microbiology and Biotechnology, 74(1), 13-21. https://doi.org/10.1007/s00253-006-0748-0 Van Hamme, J. D., Bottos, E. M., Bilbey, N. J., & Brewer, S. E. (2013). Genomic and proteomic characterization of Gordonia sp. NB4-1Y in relation to 6: 2 fluorotelomer sulfonate biotransformation. Microbiology, 159(8), 1618-1628. https://doi.org/10.1099/mic.0.068932-0 Vermeij, P., Wietek, C., Kahnert, A., Wüest, T., & Kertesz, M. A. (1999). Genetic organization of sulphur‐controlled aryl desulphonation in Pseudomonas putida S‐313. Molecular Microbiology, 32(5), 913-926. https://doi.org/10.1046/j.1365-2958.1999.01398.x Verstraete, W., & Alexander, M. (1972). Mechanism of nitrification by Arthrobacter sp. Journal of bacteriology, 110(3), 962-967.  Verstraete, W., & Focht, D. D. (1977). Biochemical ecology of nitrification and denitrification. In Advances in Microbial Ecology (pp. 135-214). Springer, Boston, MA. Vestergren, R., & Cousins, I. T. (2009). Tracking the pathways of human exposure to perfluorocarboxylates. Environmental Science & Technology, 43(15), 5565-5575. https://doi.org/10.1021/es900228k Vestergren, R., Herzke, D., Wang, T., & Cousins, I. T. (2015). Are imported consumer products an important diffuse source of PFASs to the Norwegian 135  environment? Environmental Pollution, 198, 223-230. https://doi.org/10.1016/j.envpol.2014.12.034 Vierke, L., Möller, A., & Klitzke, S. (2014). Transport of perfluoroalkyl acids in a water-saturated sediment column investigated under near-natural conditions. Environmental Pollution, 186, 7-13. https://doi.org/10.1016/j.envpol.2013.11.011 Vogel, T. M., Criddle, C. S., & McCarty, P. L. (1987). Transformations of halogenated aliphatic compounds. Environmental Science & Technology, 21(8), 722-736. https://doi.org/10.1021/es00162a001 Wang, F., & Shih, K. (2011). Adsorption of perfluorooctanesulfonate (PFOS) and perfluorooctanoate (PFOA) on alumina: Influence of solution pH and cations. Water Research, 45(9), 2925-2930. https://doi.org/10.1016/j.watres.2011.03.007 Wang, N., Szostek, B., Folsom, P. W., Sulecki, L. M., Capka, V., Buck, R. C., & Gannon, J. T. (2005a). Aerobic biotransformation of 14C-labeled 8-2 telomer B alcohol by activated sludge from a domestic sewage treatment plant. Environmental Science & Technology, 39(2), 531-538. https://doi.org/10.1021/es049466y Wang, N., Szostek, B., Buck, R. C., Folsom, P. W., Sulecki, L. M., Capka, V., & Gannon, J. T. (2005b). Fluorotelomer alcohol biotransformationdirect evidence that perfluorinated carbon chains breakdown. Environmental Science & Technology, 39(19), 7516-7528. https://doi.org/10.1021/es0506760 Wang, N., Liu, J., Buck, R. C., Korzeniowski, S. H., Wolstenholme, B. W., Folsom, P. W., & Sulecki, L. M. (2011). 6: 2 Fluorotelomer sulfonate aerobic biotransformation in activated sludge of waste water treatment plants. Chemosphere, 82(6), 853-858. https://doi.org/10.1016/j.chemosphere.2010.11.003 Wang, N., Buck, R. C., Szostek, B., Sulecki, L. M., & Wolstenholme, B. W. (2012a). 5: 3 Polyfluorinated acid aerobic biotransformation in activated sludge via novel “one-carbon removal pathways”. Chemosphere, 87(5), 527-534. https://doi.org/10.1016/j.chemosphere.2011.12.056 Wang, L., Xu, H., Cooper, W. J., & Song, W. (2012b). Photochemical fate of beta-blockers in NOM enriched waters. Science of the Total Environment, 426, 289-295. https://doi:10.1016/j.scitotenv.2012.03.031 Wang, Z., Cousins, I. T., Scheringer, M., & Hungerbühler, K. (2013). Fluorinated alternatives to long-chain perfluoroalkyl carboxylic acids (PFCAs), perfluoroalkane sulfonic acids (PFSAs) and their potential precursors. Environment International, 60, 242-248.  https://doi.org/10.1016/j.envint.2013.08.021 Wang, Z., Cousins, I. T., Scheringer, M., Buck, R. C., & Hungerbühler, K. (2014). Global emission inventories for C4–C14 perfluoroalkyl carboxylic acid (PFCA) homologues from 1951 to 2030, Part I: production and emissions from quantifiable sources. Environment International, 70, 62-75. https://doi.org/10.1016/j.envint.2014.04.013 Wang, X., Cao, A., Zhao, G., Zhou, C., & Xu, R. (2017). Microbial community structure and diversity in a municipal solid waste landfill. Waste Management, 66, 79-87. https://doi.org/10.1016/j.wasman.2017.04.023 136  Wang, B., Yao, Y., Chen, H., Chang, S., Tian, Y., & Sun, H. (2020). Per-and polyfluoroalkyl substances and the contribution of unknown precursors and short-chain (C2–C3) perfluoroalkyl carboxylic acids at solid waste disposal facilities. Science of The Total Environment, 705, 135832. https://doi.org/10.1016/j.scitotenv.2019.135832 Ward, B. B. (2008). Nitrification in marine systems. Nitrogen in the marine environment, 2, 199-261. https://doi.org/10.1128/9781555817145.ch13 Washington, J. W., Naile, J. E., Jenkins, T. M., & Lynch, D. G. (2014). Characterizing fluorotelomer and polyfluoroalkyl substances in new and aged fluorotelomer-based polymers for degradation studies with GC/MS and LC/MS/MS. Environmental Science & Technology, 48(10), 5762-5769. https://doi.org/10.1021/es500373b Washington, J. W., & Jenkins, T. M. (2015a). Abiotic hydrolysis of fluorotelomer-based polymers as a source of perfluorocarboxylates at the global scale. Environmental Science & Technology, 49(24), 14129-14135. https://doi.org/10.1021/acs.est.5b03686 Washington, J. W., Jenkins, T. M., Rankin, K., & Naile, J. E. (2015b). Decades-scale degradation of commercial, side-chain, fluorotelomer-based polymers in soils and water. Environmental Science & Technology, 49(2), 915-923.  https://doi.org/10.1021/es504347u Washington, J. W., Jenkins, T. M., & Weber, E. J. (2015c). Identification of unsaturated and 2H polyfluorocarboxylate homologous series and their detection in environmental samples and as polymer degradation products. Environmental Science & Technology, 49(22), 13256-13263. https://doi.org/10.1021/acs.est.5b03379 Washington, J. W., Rankin, K., Libelo, E. L., Lynch, D. G., & Cyterski, M. (2019). Determining global background soil PFAS loads and the fluorotelomer-based polymer degradation rates that can account for these loads. Science of the Total Environment, 651, 2444-2449. https://doi.org/10.1016/j.scitotenv.2018.10.071 Watson, S. W., Valois, F. W., & Waterbury, J. B. (1981). The family nitrobacteraceae. In The prokaryotes (pp. 1005-1022). Springer, Berlin, Heidelberg. Weinberg, I., Dreyer, A., Ebinghaus, R., 2011. Landfills as sources of polyfluorinated compounds, polybrominated diphenyl ethers and musk fragrances to ambient air. Atmospheric Environment. 45, 935-941. https://doi.10.1016/j.atmosenv.2010.11.011. Weishaar, J. L., Aiken, G. R., Bergamaschi, B. A., Fram, M. S., Fujii, R., & Mopper, K. (2003). Evaluation of specific ultraviolet absorbance as an indicator of the chemical composition and reactivity of dissolved organic carbon. Environmental Science & Technology, 37(20), 4702-4708. https://doi.org/10.1021/es030360x Wellington Laboratories. Reference and handling guides: perfluoroalkyl compounds. Retrieved 2013 August 7, from http://www.well-labs.com/docs/pfc_reference_handling_ guide.pdf, 2012 Wenk, J., Von Gunten, U., & Canonica, S. (2011). Effect of dissolved organic matter on the transformation of contaminants induced by excited triplet states and the hydroxyl radical. Environmental Science & Technology, 45(4), 1334-1340. https://doi.org/10.1021/es102212t 137  Whittaker, R. H. (1972). Evolution and Measurement of Species Diversity. Taxon, 21, 213-251.  Witzel, K. P., & Overbeck, H. J. (1979). Heterotrophic nitrification by Arthrobacter sp.(strain 9006) as influenced by different cultural conditions, growth state and acetate metabolism. Archives of Microbiology, 122(2), 137-143. https://doi.org/10.1007/BF00411352 Woldegiorgis, A., Andersson, J., Remberger, M., Kaj, L., Ekheden, Y., Blom, L., Brorström-Lundén, E., Borgen, A., Schlabach, M., (2006). Results from the Swedish National Screening Programme 2005: Subreport 3: Perflourinated Alkylated Substances (PFAS).  Wong, J. T. F., Chen, X., Deng, W., Chai, Y., Ng, C. W. W., & Wong, M. H. (2019). Effects of biochar on bacterial communities in a newly established landfill cover topsoil. Journal of environmental management, 236, 667-673. https://doi.org/10.1016/j.jenvman.2019.02.010 Xi, R., Long, X. E., Huang, S., & Yao, H. (2017). pH rather than nitrification and urease inhibitors determines the community of ammonia oxidizers in a vegetable soil. AMB Express, 7(1), 129.  https://doi.org/10.1186/s13568-017-0426-x Xiao, F. (2017). Emerging poly-and perfluoroalkyl substances in the aquatic environment: a review of current literature. Water research, 124, 482-495. https://doi.org/10.1016/j.watres.2017.07.024 Xie, Z., & Guan, W. (2015). Research on fluorescence spectroscopy characteristics of dissolved organic matter of landfill leachate in the rear part of three gorges reservoir. Journal of Spectroscopy, 2015. http://dx.doi.org/10.1155/2015/785406 Xie, S., Wang, T., Liu, S., Jones, K. C., Sweetman, A. J., & Lu, Y. (2013). Industrial source identification and emission estimation of perfluorooctane sulfonate in China. Environment International, 52, 1-8. https://doi.org/10.1016/j.envint.2012.11.004 Xing, Z., Zhao, T., Gao, Y., He, Z., Zhang, L., Peng, X., & Song, L. (2017). Real-time monitoring of methane oxidation in a simulated landfill cover soil and MiSeq pyrosequencing analysis of the related bacterial community structure. Waste Management, 68, 369-377. https://doi.org/10.1016/j.wasman.2017.05.007 Xu, F., & Bhandari, A. (2003). Retention and distribution of 1-naphthol and naphthol polymerization products on surface soils. Journal of Environmental Engineering, 129(11), 1041-1050. https://doi.org/10.1061/(ASCE)0733-9372(2003)129:11(1041) Xu, H., Cooper, W. J., Jung, J., & Song, W. (2011). Photosensitized degradation of amoxicillin in natural organic matter isolate solutions. Water Research, 45(2), 632-638. https://doi.org/10.1016/j.watres.2010.08.024 Xu, S., Lu, W., Liu, Y., Ming, Z., Liu, Y., Meng, R., & Wang, H. (2017). Structure and diversity of bacterial communities in two large sanitary landfills in China as revealed by high-throughput sequencing (MiSeq). Waste Management, 63, 41-48. https://doi.org/10.1016/j.wasman.2016.07.047 Yamazaki, E., Falandysz, J., Taniyasu, S., Hui, G., Jurkiewicz, G., Yamashita, N., & Lam, P. K. (2016). Perfluorinated carboxylic and sulphonic acids in surface water media from the regions of Tibetan Plateau: Indirect evidence on photochemical degradation?. Journal of 138  Environmental Science and Health, Part A, 51(1), 63-69. https://doi.org/10.1080/10934529.2015.1079113 Yan, H., Cousins, I. T., Zhang, C., & Zhou, Q. (2015). Perfluoroalkyl acids in municipal landfill leachates from China: Occurrence, fate during leachate treatment and potential impact on groundwater. Science of the Total Environment, 524, 23-31.  https://doi.org/10.1016/j.scitotenv.2015.03.111 Yang, S., & Song, L. (2019). Succession of bacterial community structure and metabolic function during solid waste decomposition. Bioresource Technology, 291, 121865. https://doi.org/10.1016/j.biortech.2019.121865 Yang, S., Xu, F., Wu, F., Wang, S., & Zheng, B. (2014). Development of PFOS and PFOA criteria for the protection of freshwater aquatic life in China. Science of the Total Environment, 470, 677-683. https://doi.org/10.1016/j.scitotenv.2013.09.094 Yang, X., Huang, J., Zhang, K., Yu, G., Deng, S., & Wang, B. (2014). Stability of 6: 2 fluorotelomer sulfonate in advanced oxidation processes: degradation kinetics and pathway. Environmental Science and Pollution Research, 21(6), 4634-4642. https://doi.org/10.1007/s11356-013-2389-z Yao, Y., Zhu, H., Li, B., Hu, H., Zhang, T., Yamazaki, E., & Sun, H. (2014). Distribution and primary source analysis of per-and poly-fluoroalkyl substances with different chain lengths in surface and groundwater in two cities, North China. Ecotoxicology and environmental safety, 108, 318-328. https://doi.org/10.1016/j.ecoenv.2014.07.021 Ye, F., Zushi, Y., & Masunaga, S. (2015). Survey of perfluoroalkyl acids (PFAAs) and their precursors present in Japanese consumer products. Chemosphere, 127, 262-268. https://doi.org/10.1016/j.chemosphere.2015.02.026 Yin, T., Chen, H., Reinhard, M., Yi, X., He, Y., & Gin, K. Y. H. (2017). Perfluoroalkyl and polyfluoroalkyl substances removal in a full-scale tropical constructed wetland system treating landfill leachate. Water Research, 125, 418-426. https://doi.org/10.1016/j.watres.2017.08.071 You, C., Jia, C., & Pan, G. (2010). Effect of salinity and sediment characteristics on the sorption and desorption of perfluorooctane sulfonate at sediment-water interface. Environmental Pollution, 158(5), 1343-1347. https://doi.org/10.1016/j.envpol.2010.01.009 Yu, Q., Zhang, R., Deng, S., Huang, J., & Yu, G. (2009). Sorption of perfluorooctane sulfonate and perfluorooctanoate on activated carbons and resin: kinetic and isotherm study. Water Research, 43(4), 1150-1158. https://doi.org/10.1016/j.watres.2008.12.001 Yu, X., Takabe, Y., Yamamoto, K., Matsumura, C., & Nishimura, F. (2016). Biotransformationproperty of 8: 2 fluorotelomer alcohol (8: 2 FTOH) under aerobic/anoxic/anaerobic conditions. Journal of Water and Environment Technology, 14(3), 177-190. https://doi.org/10.2965/jwet.15-056 Yu, X., Nishimura, F., & Hidaka, T. (2018a). Enhanced generation of perfluoroalkyl carboxylic acids (PFCAs) from fluorotelomer alcohols (FTOHs) via ammonia-oxidation process. Chemosphere, 198, 311-319. https://doi.org/10.1016/j.chemosphere.2018.01.132 139  Yu, X., Nishimura, F., & Hidaka, T. (2018b). Effects of microbial activity on perfluorinated carboxylic acids (PFCAs) generation during aerobic biotransformation of fluorotelomer alcohols in activated sludge. Science of the Total Environment, 610, 776-785. https://doi.org/10.1016/j.scitotenv.2017.08.075 Zainun, M. Y., & Simarani, K. (2018). Metagenomics profiling for assessing microbial diversity in both active and closed landfills. Science of the Total Environment, 616, 269-278. https://doi.org/10.1016/j.scitotenv.2017.10.266 Zeng, Y., Baumbach, J., Barbosa, E. G. V., Azevedo, V., Zhang, C., & Koblížek, M. (2016). Metagenomic evidence for the presence of phototrophic G emmatimonadetes bacteria in diverse environments. Environmental Microbiology Reports, 8(1), 139-149. https://doi.org/10.1111/1758-2229.12363 Zhang, Q. Q., Tian, B. H., Zhang, X., Ghulam, A., Fang, C. R., & He, R. (2013). Investigation on characteristics of leachate and concentrated leachate in three landfill leachate treatment plants. Waste Management, 33(11), 2277-2286. http://dx.doi.org/10.1016/j.wasman.2013.07.021 Zhang, D., Yan, S., & Song, W. (2014). Photochemically induced formation of reactive oxygen species (ROS) from effluent organic matter. Environmental Science & Technology, 48(21), 12645-12653. https://doi.org/10.1021/es5028663 Zhang, S., Lu, X., Wang, N., & Buck, R. C. (2016). Biotransformation potential of 6: 2 fluorotelomer sulfonate (6: 2 FTSA) in aerobic and anaerobic sediment. Chemosphere, 154, 224-230. https://doi.org/10.1016/j.chemosphere.2016.03.062 Zhang, S., Merino, N., Wang, N., Ruan, T., & Lu, X. (2017). Impact of 6: 2 fluorotelomer alcohol aerobic biotransformation on a sediment microbial community. Science of The Total Environment, 575, 1361-1368. https://doi.org/10.1016/j.scitotenv.2016.09.214 Zhao, L., McCausland, P. K., Folsom, P. W., Wolstenholme, B. W., Sun, H., Wang, N., & Buck, R. C. (2013a). 6: 2 Fluorotelomer alcohol aerobic biotransformation in activated sludge from two domestic wastewater treatment plants. Chemosphere, 92(4), 464. https://doi.org/10.1016/j.chemosphere.2013.02.032 Zhao, L., Folsom, P. W., Wolstenholme, B. W., Sun, H., Wang, N., & Buck, R. C. (2013b). 6: 2 Fluorotelomer alcohol biotransformation in an aerobic river sediment system. Chemosphere, 90(2), 203-209. https://doi.org/10.1016/j.chemosphere.2012.06.035      140  Appendices             Appendix A  Data on occurrence of PFASs in landfill leachate (Chapter 2)   Occurrence of perfluoroalkyl acids (PFAAs) in landfill leachate (concentrations expressed in ng/L) Class of PFASs   Kallenborn et al., (2004) Bossi et al., (2008) Woldegiorgis et al., (2008) Huset et al., (2011) Benskin et al., (2012) Busch et al., (2010) Perkola et al., (2013) Allred et al., (2014) Yan et al., (2015) Gallen et al., (2016) Fuertes et al., (2017) Country Nordic Denmark Sweden USA Canada Germany Baltic Sea USA China Australia Spain Perfluoroalkane sulfonic acids (PFSA) PFBS 5–110 (50) a -b <0.5–110 (40) 280-890 (570) 40-190 (90) <0.5–1350 (220) - 40-3200 (200) 1600-41600 (9240) <.5-840 (250) <60-580 PFHxS 10–140 (80) <0.5c–3 10–1800 (520) 160-700 (280) 85-570 (330) <0.5–180 (20) - 40-1100 (650) 30-480 (140) <.5 - 1900 (380) <30 PFOS 30–190 (80) <1–4 30–1500 (550) 60-160 (100) 220-4400 (390) 0.5–235 (30) 90-140 (110) 25-590 (150) 1150-6020 (1740) <.5 - 1100 (310) <30 PFDS - - <1 <.5-5 (1) 2-60 NDd - 10 - <.5-3 <80 Perfluoroalkyl carboxylic acids (PFCAs)   PFBA - - <12–30 (10) 170-1700 (490) 120-660 (260) <3–2970 (460) - 70-3700 (1150) 1100-9270 (2720) <.5-1600 (250) <1-790 (80) PFPeA - - - 120-1500 (610) 570-1800 (100) - - 50-3200 (1450) 610-6530 (1670) - 20-330 (200) PFHxA 26–700 (230) - <7–310 (80) 270-790 (390) 670-2500 (1300) <0.5–2510 (230) 50- 200 (120) 190-8900 (1750) 140-4430 (270) 12 - 5700 (970) 100-840 (250) PFHpA - - <20–260 (200) 100-340 (170) 240-690 (440) <0.5–280 (50) - 60-3100 (1020) 70-5830 (180) 2-3500 (540) <17-100 (60) PFOA 90–510 (290) <2–6 (3) 40–1000 (540) 380-1000 (600) 300-1500 (530) <0.5–920 (150) 75- 270 (170) 150-5000 (1050) 280-214000 (2260) 19-2100 (450) 200-510 (440) PFNA 5–60 (30) <1 <20–100 (40) 20-30 (20) 30-450 (60) <3–80 (10) - 10-290 (20) 1-380 (170) <.5-90 (20) <50 PFDA - <1 <20–220 (80) 0.5-23 (15) 40-1100 (110) <0.5–55 (5) 2- 4 (2) 6-200 (10) 1-19 (10) <.5-57 (10) <53 PFUnA - - - 0-10 4-120 (10) <0.5–3 - <0.5 - <.5-20 (5) <28 arange of concentration (median value); bnot analyzed; c‘<’ represents detection limit; dND: not detected;142   Occurrence of perfluoroalkyl sulfonamide derivatives, fluorotelomer acids in landfill leachate (concentration expressed as ng/L) Class of PFASs   Huset et al., (2011) Benskin et al., (2012) Busch et al., (2012) Allred et al., (2014) Country USA Canada Germany USA Landfill sites  6 1 22 6 Perfluoroalkane sulfonamidoacetic acids (FASAAs) and N-alkylPerfluoroalkane sulfonamidoacetic acidsa  FOSA 0-7 (1) 5-90 (10)d <0.5–14.0 (3) - FOSAA 0-1 (1) 20-990 (47) - - MeFBSAA 60-440 (160) - - 40-2900 (720) MeFPeSAA  - - - 20-660 (400) MeFHxSAA - - - 15-1900 (190) MeFHpSAA  - - - 2 -140 (10) MeFOSAA 15-290 (140) 30-5000 (350) - 20-990 (80) EtFBSAA  - - - 5-100 (30) EtFPeSAA  - - - 2-50 (30) EtFHxSAA  - - - 4-50 (20) EtFHpSAA  - - - 10-15 (10) EtFOSAA 20-480 (90) 290-8700 (590) - 10-310 (70) n:2 Fluorotelomer carboxylic acids (n:2 FTCAs) and unsaturated carboxylic acidsb (n:2 FTUCAs) 6:2 FTCA - 40-280 (120) - 230-2000 (550) 8:2 FTCA - 190-5200 (360) - 20-240 (150) 10:2 FTCA - 30-770 (80) - 15 3:3 FTCA  - - - 8-55 (40) 5:3 FTCA  - - - 320-18000 (3600) 7:3 FTCA - - - 20-1700 (200) 6:2 FTUCA - 5-65 (10) - 20 8:2 FTUCA - 50-2100 (130) - 2 10:2 FTUCA - 5-430 (30) - - n:2 Fluorotelomer sulfonic acidsc (n:2 FTSs) 4:2 FTS -   - 5-12 (6) 6:2 FTS 30-370 (30)   - 20-470 (120) 8:2 FTS 10-120 (30)   - 5-150 (80) adegradation intermediates of (N-ethyl/methyl) perfluoroalkane sulfonamidethanol;  bdegradation intermediate of fluorotelomer alcohols; d: range of concentration (median)             143   Classification of landfill leachate according to age and typical characteristics (Renou et al., 2008)   Young Intermediate Mature Age (years) <5 5 – 10 >10 pH-value 6.5 6.5 – 7.5 >7.5 CODa (g/L) >10 4 – 10 <4 BODb5/COD >0.3 0.1 – 0.3 <0.1 Organic composition 80% volatile fatty acids  5 – 30% volatile fatty acids + humic and fulvic acids humic and fulvic acids      aCOD: chemical oxygen demand; bBOD5:5-day biochemical oxygen demand  144   Appendix B  Supplemental Information on Aerobic Biotransformation of Fluorotelomer Compounds in Landfill Leachate-Sediment System (Chapter 3)  Standards of poly- and perfluoroalkyl substances (PFASs)  Analyte Acronym Formula Chemical Structure Accurate Mass CAS Source 2-Perfluorohexyl ethanol 6:2 FTOH C6F13CH2CH2OH  364.1 647-42-7 Wellington Laboratories, Canada 2-Perfluorooctyl ethanol 8:2 FTOH C8F17CH2CH2OH  464.1 678-39-7 Wellington Laboratories, Canada 1-Perfluoroheptyl ethanol  7:2 sFTOH C7F15CH(OH)CH3  414.1 24015-83-7 Wellington Laboratories, Canada 1-Perfluoropentyl ethanol 5:2 sFTOH C5F11CH(OH)CH3  314.1 914637-05-1 Wellington Laboratories, Canada Perfluoro-n-butanoic acid PFBA C3F7COOH  214.03 377-22-4 Wellington Laboratories, Canada Perfluoro-n-pentanoic acid PFPeA C4F9COOH  264.04 2706-90-3 Wellington Laboratories, Canada Perfluoro-n-hexanoic acid PFHxA C5F11COOH  314.05 307-24-4 Wellington Laboratories, Canada 145  Analyte Acronym Formula Chemical Structure Accurate Mass CAS Source Perfluoro-n-heptanoic acid PFHpA C6F13COOH  364.06 375-85-9 Wellington Laboratories, Canada Perfluoro-n-octanoic acid PFOA C7F15COOH  414.07 335-67-1 Wellington Laboratories, Canada Perfluoro-n-nonanoic acid PFNA C8F17COOH  464.07 375-95-1 Wellington Laboratories, Canada 2H-Perfluoro-2-octenoic acid  6:2 FTUCA C8H2F12O2  358.08 70887-88-6 Wellington Laboratories, Canada 2H-Perfluoro-2-decenoic acid 8:2 FTUCA C10H2F16O2  458.09 70887-84-2 Wellington Laboratories, Canada 2-Perfluorohexyl ethanoic acid 6:2 FTCA C8H3F13O2  378.08 53826-12-3 Wellington Laboratories, Canada 2-Perfluorooctyl ethanoic acid 8:2 FTCA C10H3F17O2  478.1 27854-31-5 Wellington Laboratories, Canada 3-Perfluoropentyl propanoic acid  5:3 FTCA C8H5F11O2  342.1 914637-49-3 Wellington Laboratories, Canada 3-Perfluoroheptyl propanoic acid 7:3 FTCA C10H5F15O2  442.1 812-70-4 Wellington Laboratories, Canada 146  Analyte Acronym Formula Chemical Structure Accurate Mass CAS Source Sodium 1H,1H,2H,2H-perfluorooctane sulfonate Na salt of 6:2 FTS C8H4F13SO3Na C C C C C C CH2 CH2FFFF FFFF FF FFFSOOONa  450.1 27619-97-2 Synquest Laboratories, Florida, USA,  Wellington Laboratories, Canada Sodium 1H,1H,2H,2H-perfluoro-1-[1,2-13C2]- octane sulfonate (6:2) (M+2) 6:2 FTS - C C C C C C CH2 CH2FFFF FFFF FF FFFSOOONa13 13 452.1 N/A Wellington Laboratories, Canada 2H-Perfluoro-[1,2-13C2]-2-decenoic acid (M+2) 8:2 FTUCA - C C C C C C CFFFFF FF FF FFFF FFCCH13C13FOHO(M+2)_8:2 FTUCA 460.09 N/A Wellington Laboratories, Canada 2-Perfluorooctyl-[1,1-2H2]-[1,2-13C2]-ethanol (M+4) 8:2 FTOH - C C C C C C C C13C13FFF FFFFFFFF FFFFFFHHDOHD8:2 FTOH(M+4) 468.1 N/A Wellington Laboratories, Canada Perfluoro-n-[2,3,4-13C3] butanoic acid (M+3) PFBA - C13FFC13C13CFFFF OOHF(M+3)PFBA 218 N/A Wellington Laboratories, Canada Perfluoro-n-[1,2,3,4,6-13C5] hexanoic acid (M+5) PFHxA - C13C C13FFFFFFC13C13C13FFFF OOHF(M+5) PFHxA 319 N/A Wellington Laboratories, Canada Perfluoro-n-[13C8] octanoic acid (M+8) PFOA - C13C C13FFFFFFC13C13C13FFFF OOHF(M+5) PFHxAC13C13C13C13C13FFFF FF FFFFC13C13C13FFFF OOHFPFOA(M+8) 422 N/A Wellington Laboratories, Canada    147   Figure B.1 Microcosm bottles containing landfill leachate and sediment    Organic and inorganic contents analysis The ammonia, nitrate and nitrite were analyzed by a QuickChem® (Lachat Instrumets, WI, USA) automated ion analyzer. A Shimadzu total organic carbon (TOC) analyzer (TOC-LCSH/CPN) was used for TOC analysis of the dissolved phase of the microcosms. Anions were analyzed by a Dionex™ (ICS-900) ion chromatography (IC) (Dionex Inc., Sunnyvale, CA, USA) system. Anions in IC were separated using an IonPac™ AS4A-SC (4 mm × 250 mm) analytical column and IonPac™ AG4A-SC guard column. The mobile phase consisted of 3.5 mM sodium carbonate and 1.0 mM sodium bicarbonate (flow rate of 1.2 mL/min). A total of 24 metals were analyzed using an inductively coupled plasma spectrometer (ICP), with a detection limit of 0.5 mg/L. The metals included Aluminum (Al), Arsenic (As), Boron (B), Barium (Ba), Beryllium (Be), Calcium (Ca), Cadmium (Cd), Cobalt (Co), Copper (Cu), Chromium (Cr), Iron (Fe), Magnesium (Mg), Manganese (Mn), Molybdenum (Mo), Nickel (Ni), Phosphorus (P), Silver (Ag), Antimony (Sb), Selenium (Se), Silicon (Si), Tin (Ti), Thallium (Tl), Vanadium (V), Zinc (Zn).  148      List of chemicals used for mineral media preparation  Chemicals  Supplier Purity Potassium phosphate, Monobasic (KH2PO4) Millipore Sigma, Canada ≥ 99% Sodium phosphate dibasic dihydrate (Na2HPO4·2H2O) Fisher chemical, USA ≥ 98% Ammonium chloride (NH4Cl) VWR Life Science  ≥99%  Calcium chloride dihydrate (CaCl2·2H2O) Sigma-Aldrich, USA ≥99% Magnesium Sulfate Heptahydrate (MgSO4·7H2O) Fisher chemical, USA ≥98% Iron (III) chloride hexahydrate (FeCl3·6H2O) Millipore Sigma, Canada ≥98%         149       Instrumental method parameters for analysis of PFCAs and Fluorotelomer acids by LC-MS/MS  Instruments Agilent Technologies 1200 liquid chromatography 6410 Triple Quad mass spectrometer. Operated in the negative ion multiple reaction monitoring mode. Analytical Column Waters Xterra C18MS analytical column, 100 mm length, 2.1 mm ID, 3.5 μm particle size;  Guard Column Waters Xterra C18MS analytical column, 30 mm length, 2.1 mm ID, 3.5 μm particle size LC conditions Column temperature (oC): 50 Maximum Pressure (bar): 345 Flow rate (mL/min): 0.7 Mobile Phases A: 10 mM Ammonium acetate and 10 mM acetic acid in water B: Acetonitrile Gradient Profile Time (min) A (%) B (%) 0 80 20 6 40 60 7 10 90 8 10 90 9 80 20 12 80 20  MS conditions Source temperature (oC): 120 Desolvation temperature (oC): 325 Capillary voltage (kV): 3.5 Injection Volume: 4 μL Monitored Ion Transitions Analytes Fragmentor voltage (V) Collision energy (eV) Parent to product ion transition PFBA 28 0 213.0 -> 169.0 PFPeA 20 0 263.0 -> 219.0 PFHxA 30 0 313.0 -> 269.0 PFHpA 30 0 363.0 -> 319.0 PFOA 56 2 413.0 -> 369.0 PFNA 30 2 463.0 -> 419.0 5_3 FTCA 70 2 341.0 -> 237.0 7_3 FTCA 70 2 377.0 -> 293.0 6_2 FTUCA 70 2 357.0 -> 293.0 8_2 FTUCA 70 2 457.0 -> 393.0 6_2 FTS 130 2 427.0 -> 407.0 M3FPBA 28 0 216.0-> 172.0 150  M2_6:2 FTS 130 2 429.0 -> 409.0 M5PFHxA 30 0 318.0 -> 273.0 M2_8:2 FTUCA 75 18 459.0 -> 394.0 M8PFOA 56 2 413.0 -> 369.0  Calibrations  The target compounds were quantified using a 7-point calibration in the working concentration range (2 to 500 ng/mL) with 13C3 PFBA and 13C5 PFHxA as injection internal standards. The recoveries of labelled-surrogate compound were used to correct recovery values for all target compounds. The limits of detection (LODs) and limits of quantification (LOQs) (Table B6) were calculated using based on signal-to-noise ratios of 3 and 10, respectively, determined at the lowest method calibration limit (2 ng/mL).   Instrumental method parameters for analysis of FTOHs by GC-MS     Instrument: Agilent 7890B gas chromatograph interfaced 5977B mass spectrometer operated in the chemical ionization mode Analytical Column: Agilent CP-Sil 8 CB (30 m, 0.25 mm, 0.25 µm, 7-inch cage) Injection Volume: 1 μL Monitored Ion Transitions: Analytes Ion Transitions (m/z) 7:2s FTOH 415 > 377 5:2s FTOH 315>277 8:2 FTOH 465 > 427 6:2 FTOH 365>327 [M+4] 8:2 FTOH 469 > 431  GC Parameters: GC system inlet temperature: 250°C Column temperature: held at 60°C for 1 min, and then ramped up at 3°C/min to 75°C, then at 20°C/min to 185°C with ballistic heating to a final temperature of 260°C, which was held for 6 min Mode: Pulsed spitless MS Parameters: MS source temperature: 300 ◦C Quadrupoles temperature: 150 ◦C Calibration  Quantitation was achieved with a 6-point linear regressed calibration curve spanning 5 to 500 ng/ml. 151   Detection limits of target analytes in injection solvent (80:20/Water:Acetonitrile,v/v) Analytes Internal Standard  Surrogate Standard Limit of detection (LOD)a (µg/L) Limit of Quantification (LOQ)b (µg/L)  6:2 FTOH M4_8:2 FTOH N/AC 1.4 4.8 8:2 FTOH M4_8:2 FTOH N/A 1.0 3.2 7:2 sFTOH M4_8:2 FTOH N/A 0.5 1.7 5:2 sFTOH M4_8:2 FTOH N/A 0.8 2.6 PFBA M3PFBA M8PFOA 1.0 2.9 PFPeA M3PFBA M8PFOA 0.9 2.6 PFHxA M5PFHxA M8PFOA 0.3 1.0 PFHpA M5PFHxA M8PFOA 0.2 0.6 PFOA M5PFHxA M8PFOA 0.2 0.5 PFNA M5PFHxA M2_8:2 FTUCA 0.3 1 6:2 FTUCA M5PFHxA M2_8:2 FTUCA 0.3 1.0 8:2 FTUCA M5PFHxA M2_8:2 FTUCA 0.9 2.6 6:2 FTCA M5PFHxA M2_8:2 FTUCA 1 3 8:2 FTCA M5PFHxA M2_8:2 FTUCA 0.6 1.8 5:3 FTCA M5PFHxA M2_8:2 FTUCA 0.3 1.0 7:3 FTCA M5PFHxA M2_8:2 FTUCA 0.1 0.3 6:2 FTS M5PFHxA M2FTS 0.5 1.5 (M+2) 6:2 FTS M5PFHxA N/A 0.7 2.2 (M+2) 8:2 FTUCA N/A N/A 0.2 0.6 (M+4) 8:2 FTOH N/A N/A 1.1 3.6 (M+3) PFBA N/A N/A 0.02 0.07  (M+5) PFHxA N/A N/A 0.01 0.03 (M+8) PFOA N/A N/A 0.1 0.3 asignal to noise ratio of 3 determined at lowest calibration; bcalculated as 3×LOD; CN//A: not applicable          152   Figure B.2 Evolution of 7:3 FTUCA (m/z: 439>369) over time (days) in 8:2 FTOH spiked live samples. Due to lack of authentic standard, concentrations of 7:3 FTUCA has been approximated by the area counts of mass spectrometry peak of sediment-leachate extracts.    Figure B.3 Plots showing variation natural logarithm of the remaining parent compound concentrations for a) 8:2 FTOH and b) 6:2 FTS. the trend lines were used for estimating the pseudo-first order kinetics and half-lives of 8:2 FTOH and 6:2 FTS in landfill leachate-sediment.     0204060801001201400 20 40 60 80 100Peak Area Count (Arbitratry unit)Time (days)y = 1.1497e-0.008xR² = 0.864800.40.81.21.60 20 40 60 80 100ln (concentration)Time (days)b)y = 6.2133e-0.001xR² = 0.77955.45.65.866.26.40 20 40 60 80 100ln (concentration)Time (days)a)153    Physical and chemical monitoring data for live-control microcosms (plotted in Figure 3.1)   Sampling Day 0 1 3 7 14 30 45 60 90 pH  Replicate 1 8.0 7.9 7.2 7.2 7.1 6.9 7.0 7.1 7.0 Replicate 2 7.9 7.8 7.1 7.1 7.1 7.0 7.0 7.7 7.1 Average 8.0 7.9 7.2 7.1 7.1 7.0 7.0 7.4 7.1 Error 0.0 0.0 0.0 0.0 0.0 0.1 0.0 0.3 0.0 TOC (mg/L) Replicate 1 205.9 195.7 190.5 131.3 157.9 147.7 169.3 137.5 138.0 Replicate 2 203.8 201.5 163.1 215.6 175.4 179.5 139.0 170.1 128.5 Average 204.9 198.6 176.8 173.5 166.7 163.6 154.2 153.8 133.3 Error 1.1 2.9 13.7 42.2 8.8 15.9 15.2 16.3 4.8 Ammonia (mg/L) Replicate 1 57.1 50.6 29.9 21.9 22.9 10.3 0.0 0.0 0.0 Replicate 2 62.4 50.4 26.5 20.7 10.8 0.0 0.0 0.0 0.0 Average 59.7 50.5 28.2 21.3 16.9 5.2 0.0 0.0 0.0 Error 2.7 0.1 1.7 0.6 6.0 5.2 0.0 0.0 0.0 Nitrate (mg/L) Replicate 1 3.1 3.1 22.6 44.2 55.6 57.2 60.4 65.4 64.6 Replicate 2 3.3 2.9 27.0 44.4 52.8 62.0 59.6 47.6 67.8 Average 3.2 3.0 24.8 44.3 54.2 59.6 60.0 56.5 66.2 Error 0.1 0.1 2.2 0.1 1.4 2.4 0.4 8.9 1.6 Headspace oxygen (%) Replicate 1 16.5 16.4 16.2 14.9 12.1 5.4/16.4a 14.0 11.1 5.2 Replicate 2 16.5 16.5 16.1 14.9 12.3 7.4/16.4a 11.0 10.2 6.0 Average 16.5 16.5 16.2 14.9 12.2 6.4/16.4a 12.5 10.7 5.6 Error 0.0 0.1 0.0 0.0 0.1 1 1.5 0.4 0.4 aheadspace oxygen content following reaeration           154    Concentration (nmol/L) of biotransformation products of 8:2 FTOH in live-spiked microcosms (plotted in Figure 3.3)  Sampling Day 0 1 3 7 14 30 45 60 90 PFHxA Replicate 1 1.8 1.7 2.2 3.6 2.4 2.1 2.3 4.4 3.1 Replicate 2 1.4 2.7 1.9 3.6 1.7 2.2 2.3 1.5 2.5 Average  1.6 2.2 2.0 3.6 2.0 2.2 2.3 2.9 2.8 absolute error 0.2 0.5 0.2 0.0 0.3 0.1 0.0 1.5 0.3 PFHpA Replicate 1 2.7 1.0 1.5 2.4 4.8 2.0 1.8 4.5 2.7 Replicate 2 0.7 1.6 1.1 2.4 1.1 1.7 2.2 1.1 3.2 Average  1.7 1.3 1.3 2.4 2.9 1.8 2.0 2.8 3.0 absolute error 1.0 0.3 0.2 0.0 1.9 0.1 0.2 1.7 0.3 PFOA Replicate 1 8.5 3.0 5.3 10.7 13.9 12.3 8.2 18.8 14.9 Replicate 2 2.5 4.1 5.3 10.7 6.7 11.2 13.8 11.4 14.8 Average  5.5 3.6 5.3 10.7 10.3 11.8 11.0 15.1 14.8 absolute error 3.0 0.6 0.0 0.0 3.6 0.6 2.8 3.7 0.1 8:2 FTUCA Replicate 1 2.0 0.1 0.4 0.8 1.0 1.2 1.4 10.6 6.0 Replicate 2 0.1 0.2 0.5 0.8 0.3 1.8 1.2 3.8 5.5 Average  1.1 0.2 0.5 0.8 0.7 1.5 1.3 7.2 5.7 absolute error 1.0 0.0 0.1 0.0 0.4 0.3 0.1 3.4 0.3 7:3 FTCA Replicate 1 0.7 0.9 0.8 0.6 0.9 1.6 1.9 2.6 6.5 Replicate 2 0.7 1.0 1.1 0.6 1.0 2.7 1.1 1.5 8.2 Average  0.7 0.9 0.9 0.6 0.9 2.1 1.5 2.1 7.4 absolute error 0.0 0.0 0.1 0.0 0.0 0.5 0.4 0.6 0.9 7:2 sFTOH Replicate 1 0.0 0.7 0.0 0.0 0.0 0.0 1.0 0.5 0.0 Replicate 2 0.0 0.4 0.0 0.0 0.0 0.0 0.3 0.8 0.0 Average  0.0 0.5 0.0 0.0 0.0 0.0 0.7 0.6 0.0 absolute error 0.0 0.1 0.0 0.0 0.0 0.0 0.4 0.1 0.0 7:3 FTUCA (tentative area count) Replicate 1 20.2 13.4 25.2 20.0 31.9 89.0 100.6 68.1 0.0 Replicate 2 0.0 20.2 20.2 20.0 27.2 139.6 69.0 8.4 0.0 Average  10.1 16.8 22.7 20.0 29.6 114.3 84.8 38.2 0.0 absolute error 10.1 3.4 2.5 0.0 2.4 25.3 15.8 29.8 0.0    155   Concentrations (nmol/L) of 8:2 FTOH in the live-spiked microcosms (plotted in Figure 3.4)  Sampling day 0 1 3 7 14 30 45 60 90 Solids extract Replicate 1 508.0 481.2 372.4 423.7 336.4 312.9 140.1 228.0 208.7 Replicate 2 533.6 471.1 286.9 423.7 364.3 313.7 208.6 250.1 228.4 Average 520.8 476.1 329.7 423.7 350.3 313.3 174.3 239.1 218.5 Absolute error 12.8 5.1 42.7 0.0 14.0 0.4 34.3 11.0 9.8 Headspace Replicate 1 0.0 79.6 66.8 92.3 109.8 92.4 122.5 129.0 150.0 Replicate 2 0.0 70.2 107.5 95.7 109.6 90.7 108.6 117.7 105.6 Average 0.0 74.9 87.1 94.0 109.7 91.5 115.5 123.3 127.8 Absolute error 0.0 4.7 20.4 1.7 0.1 0.8 7.0 5.7 22.2 Total 8:2 FTOH Replicate 1 508.0 560.8 439.2 516.0 446.1 405.2 262.6 357.1 358.8 Replicate 2 533.6 541.3 394.4 519.4 473.9 404.4 317.2 367.8 334.0 Average 520.8 551.1 416.8 517.7 460.0 404.8 289.9 362.4 346.4 Absolute error 12.8 9.8 22.4 1.7 13.9 0.4 27.3 5.4 12.4   Concentrations (nmol/L) of 8:2 FTOH in the sterile-spiked microcosms (plotted in Figure 3.4)  Sampling day 0 1 3 7 14 45 60 90 Solids extract Replicate 1 439.5 303.0 366.5 382.9 247.5 261.8 267.7 303.7 Replicate 2 439.5 372.7 339.5 321.4 267.3 391.0 316.6 303.7 Average 439.5 337.8 353.0 352.2 257.4 326.4 292.1 303.7 Absolute error 0.0 34.8 13.5 30.7 9.9 64.6 24.5 0.0 Headspace Replicate 1 0.0 58.2 79.5 71.1 71.9 129.3 106.3 98.1 Replicate 2 0.0 78.7 63.5 72.0 128.3 93.2 107.8 106.6 Average 0.0 68.4 71.5 71.6 100.1 111.2 107.0 102.4 Absolute error 0.0 2.0 1.5 0.1 5.4 3.5 0.1 0.8 Total 8:2 FTOH Replicate 1 439.5 361.2 446.1 454.1 319.4 391.1 374.0 401.8 Replicate 2 439.5 451.4 403.0 393.4 395.5 484.2 424.4 410.3 Average 439.5 406.3 424.6 423.7 357.5 437.7 399.2 406.0 Absolute error 0.0 45.1 21.5 30.3 38.1 46.6 25.2 4.2     156   Concentration (nmol/L) of PFASs in live control microcosms (plotted in Figure 3.2)  Sampling day 0 14 60 90 PFBA Replicate 1 0.0 0.0 0.0 0.0 Replicate 2 0.0 0.0 0.0 0.0 Average  0.0 0.0 0.0 0.0 Error 0.0 0.0 0.0 0.0 PFPeA Replicate 1 1.2 2.2 1.4 1.2 Replicate 2 1.6 2.9 1.1 1.0 Average  1.4 2.5 1.3 1.1 Error 0.2 0.4 0.1 0.1 PFHxA Replicate 1 1.7 2.2 2.0 1.3 Replicate 2 1.6 2.0 1.9 1.4 Average  1.6 2.1 1.9 1.3 Error 0.1 0.1 0.0 0.0 PFHpA Replicate 1 0.9 1.4 1.2 0.5 Replicate 2 1.1 1.6 1.0 1.0 Average  1.0 1.5 1.1 0.7 Error 0.1 0.1 0.1 0.2 PFOA Replicate 1 2.1 1.3 2.2 1.7 Replicate 2 1.9 1.5 2.1 3.2 Average  2.0 1.4 2.2 2.4 Error 0.1 0.1 0.0 0.8 5:3 FTCA Replicate 1 0.4 0.3 1.1 0.3 Replicate 2 0.5 0.3 0.0 0.3 Average  0.4 0.3 0.6 0.3 Error 0.1 0.0 0.6 0.0 7:3 FTCA Replicate 1 0.7 0.5 0.7 1.6 Replicate 2 0.9 0.7 0.5 0.3 Average  0.8 0.6 0.6 0.9 Error 0.1 0.1 0.1 0.6        157  Appendix C  Supplemental Information on Effect of Leachate Addition on 6:2 FTS Biotransformation (Chapter 4)  Standards of Perfluorocarboxylic acids (PFCAs) and 6:2 FTS and their suppliers Analyte Acronym Chemical Formula Accurate Mass CAS Source Perfluoro-n-butanoic acid PFBA C3F7COOH 213 377-22-4 Wellington Laboratories, Canada Perfluoro-n-pentanoic acid PFPeA C4F9COOH 263.9 2706-90-3 Wellington Laboratories, Canada Perfluoro-n-hexanoic acid PFHxA C5F11COOH 318.2 307-24-4 Wellington Laboratories, Canada Perfluoro-n-heptanoic acid PFHpA C6F13COOH 363.9 375-85-9 Wellington Laboratories, Canada Perfluoro-n-octanoic acid PFOA C7F15COOH 418.2 335-67-1 Wellington Laboratories, Canada Sodium 1H,1H,2H,2H-perfluorooctane sulfonate (6:2) 6:2 FTS C8H4F13SO3Na 450.1 27619-97-2 Synquest Laboratories, Florida, USA Sodium 1H,1H,2H,2H-perfluoro-1-[1,2-13C2]- octane sulfonate (6:2) M2_6:2 FTS -   Wellington Laboratories, Canada Perfluoro-n-[2,3,4-13C ]butanoic acid M3PFBA -   Wellington Laboratories, Canada Perfluoro-n-[1-13C 5]hexanoic acid MPFHxA -   Wellington Laboratories, Canada Perfluoro-n-[13C 8]octanoic acid M8PFOA -   Wellington Laboratories, Canada       158   Extraction and Clean-up of PFASs from Microcosm Extracts  Extraction of perfluoroalkyl carboxylic acids (PFCAs) For extraction of the 6:2 FTS biotransformation products, 5 ml sample was added with 50 ng of MPFHxA as surrogate standard. The sample was vortex mixed and allowed to rest for 1 hour. After adding 5 mL 50% (v/v) acetonitrile and methanol, the samples were placed on horizontal shaker (180 rpm) at 40°C for 1 day, and then centrifuged at 5000 rpm for 10 min to collect the supernatants. For the second extraction, 4.5 mL acetonitrile and 0.5 mL of 250 mM NaOH were added to the samples. The samples were again shaken at 40°C overnight, centrifuged at 5000 rpm for 10 min and the extracts were collected. The first and second extracts were mixed together, diluted to ~40 mL using D and the pH was adjusted to 6.5±0.5. The diluted extract was cleaned up by solid-phase extraction (SPE) using Oasis WAX® following a modified method (USEPA 2011). Each cartridge was first conditioned with 0.3% NH4OH in methanol, followed by 0.1 M formic acid in reagent water. Then, the cartridge was equilibrated with deionized water. After loading the sample drop-wise onto the cartridge (~5 mL/min), the cartridge was washed with 20% MeOH in 80% 0.1 M formic acid in reagent water, followed by 0.3% (v/v) NH4OH in reagent water. The cartridge was then dried by subjecting to vacuum for 5 min. Finally, the cartridge was eluted into a 15-mL clean glass centrifuge tube with 4 mL 0.3% NH4OH (v/v) in MeOH. The cleaned-up extract was dried near completeness under a gentle nitrogen stream and reconstituted with 500 µL of injection solvent (95% aqueous methanol). The reconstituted extract was spiked with internal standard (30 ng of M3PFBA and M8PFOA) prior to instrumental analysis using LC-MS/MS.  Extraction of 6:2 Fluorotelomer sulfonate (6:2 FTS) To quantify 6:2 FTS, 0.5 mL of sample was added with 150 ng of M2_6:2FTS as surrogate standard, vortex mixed and allowed to sit for one hour. Two cycles of extraction were performed using 2.5 mL of methanol and overnight shaking for 18 hours in each cycle. The liquid extracts were collected following centrifugation at 5000 rpm for 5 min and filtered through 0.45 µm PVDF syringe filter. 0.5 mL of the filtered extract was aliquoted in LC vial, added with internal standard (30 ng of M8PFOA) and analyzed using LC-MS/MS.  159   Instrumental method parameters for analysis of PFCAs and 6:2 FTS by LC-MS/MS Instruments Agilent Technologies 1200 HPLC 6430 Triple Quad mass spectrometer. Analytical Column Waters Xterra C18MS analytical column, 100 mm length, 2.1 mm ID, 3.5 μm particle size;  Guard Column Waters Xterra C18MS analytical column, 30 mm length, 2.1 mm ID, 3.5 μm particle size LC conditions Column temperature (oC): 50 Maximum Pressure (bar): 400 Mobile Phases A: Type 1 water with 5 mM ammonium acetate B: 95% aqueous methanol with 5 mM ammonium acetate  Gradient Profile Time (min) A (%) B (%) Flow Rate (ml/min) 0 85 15 0.5 0.3 85 15 0.5 2 50 50 0.6 3 35 65 0.6 5 30 70 0.6 5.10 0 100 0.6 5.60 0 100 0.6 5.70 85 15 0.5 6.70 85 15 0.5  MS conditions Gas Temperature (oC): 350°C Gas Flow: 9 l/min Nebuliser Pressure: 40 psi Acquisition mode: Negative ion MRM Capillary voltage (V): 1000 Injection Volume: 6 μL Monitored Ion Transitions Analytes Parent to product ion transition Fragmentor voltage (V) Collision energy (eV) Cell Acceleration Voltage (V) PFBA 213.0 -> 169.0 68 4 2 PFPeA 263.0 -> 219.0 52 4 2 PFHxA  313.0 -> 269.0 313.0 -> 119.0 (Qualifier) 62 62 4 20 2 2 PFHpA  363.0 -> 319.0 363.0 -> 169.0 (Qualifier) 66 66 4 16 2 2 PFOA  413.0 -> 369.0 413.0 -> 169.0 (Qualifier) 72 72 4 16 2 2 6_2 FTS  427.0 -> 407.0 427.0 -> 81.0 (Qualifier) 90 90 22 50 2 2 M3FPBA 216.0-> 172.0 64 8 2 M2_6:2 FTS  429.0 -> 409.0 429.0 -> 81 (Qualifier) 38 38 24 40 2 2 M5PFHxA 315.0 -> 270.0 62 4 2 160  M8PFOA  421.0 -> 376.0 421.0 -> 172.0 (Qualifier) 70 70 8 16 4 4  Calibrations  The target compounds were quantified using a 6-point calibration in the working concentration range (2 to 500 ng/mL) with 13C3 PFBA (M3PFBA) and 13C8 PFOA (M8PFOA) as injection internal standards. The recovery of labelled-surrogate compounds were used to correct recovery values for all target compounds. The limits of detection (LODs) and limits of quantification (LOQs) (Table S6) were calculated using based on signal-to-noise ratios of 3 and 10, respectively, determined at the lowest method calibration limit (2 ng/mL).      Detection limits (ng/mL) of target analytes in injection solvent (95% aqueous methanol) Analytes Internal Standard (ISTD) Surrogate Standard Limit of detection (LOD)a (µg/L) Limit of Quantification (LOQ)b (µg/L) PFBA M3PFBA MPFHxA 1.0 2.9 PFPeA M3PFBA MPFHxA 0.9 2.6 PFHxA M8PFOA MPFHxA 0.3 1.0 PFHpA M8PFOA MPFHxA 0.2 0.6 PFOA M8PFOA MPFHxA 0.2 0.5 6:2 FTS M8PFOA M2_6:2 FTS 0.5 1.4 M2_ 6:2 FTS M8PFOA - 0.7 2.2 M3PFBA - - 0.0 0.1 MPFHxA M8PFOA - 0.0 0.0 M8PFOA - - 0.1 0.3 asignal to noise ratio of 3 determined at lowest calibration; bcalculated as 3×LOD       161   Figure C.1 Relative abundance of microbial community structures at the phylum level in a) 6:2 FTS spiked and b) control microcosms. Lower abundance (<1%) taxa are summed and represented as “Others”. Duplicates were analyzed for 0 and 14 d samples. The absolute difference of relative abundances of duplicates samples ranged between 1 to 9%.   Relative abundance Relative abundance 162   Two-way ANOVA comparing Shannon indices of live-spiked microcosms on various sampling days under various treatment conditions  Live-spiked microcosms             Day 0 Day 7 Day 14 Day 30 Day 60  DI water 4.5 5.1 5.4 6.3 6.5  Diluted leachate 4.5 5.3 5.4 6.0 6.5  Leachate 4.8 5.3 5.5 5.9 6.3         Anova: Two-Factor Without Replication           SUMMARY Count Sum Average Variance   DI water 5 27.7 5.5 0.7   Diluted leachate 5 27.7 5.5 0.6   Leachate 5 27.7 5.5 0.3          Day 0 3 13.8 4.6 0.0   Day 7 3 15.7 5.2 0.0   Day 14 3 16.3 5.4 0.0   Day 30 3 18.1 6.0 0.0   Day 60 3 19.3 6.4 0.0                 ANOVA       Source of Variation SS df MS F P-value F crit Rows 3.41E-05 2 1.70614E-05 0.000697 0.999303 4.458970108 Columns 5.956619 4 1.489154667 60.85938 4.99E-06 3.837853355 Error 0.19575 8 0.024468778           Total 6.152403 14                 163   Two-way ANOVA comparing Shannon indices of live-control microcosms on various sampling days under various treatment conditions   Live-control microcosms             Day 0 Day 7 Day 14 Day 30 Day 60  DI water 4.5 5.2 6.0 6.0 6.6  Diluted leachate 4.5 5.3 5.3 5.8 6.5  Leachate 4.8 5.2 5.0 5.8 6.3         Anova: Two-Factor Without Replication           SUMMARY Count Sum Average Variance   DI water 5 28.3 5.7 0.6   Diluted leachate 5 27.5 5.5 0.5   Leachate 5 27.0 5.4 0.4          Day 0 3 13.8 4.6 0.0   Day 7 3 15.6 5.2 0.0   Day 14 3 16.3 5.4 0.3   Day 30 3 17.6 5.9 0.0   Day 60 3 19.4 6.5 0.0                 ANOVA       Source of Variation SS df MS F P-value F crit Rows 0.1652 2 0.0826 1.2249 0.3435 4.4590 Columns 5.7721 4 1.4430 21.4040 0.0002 3.8379 Error 0.5393 8 0.0674           Total 6.4766 14                      164  Appendix D  Supplemental Information on Role of Microbial Communities in Formation of Perfluorocarboxylic Acids from 6:2 Fluorotelomer Sulfonate (6:2 FTS) in Leachate (Chapter 5)   Figure D.1 Microcosm setup a) in Wheaton glass bottles and b) bottom of the bottles with microbial growth   Relative abundances of Actinobacteria phyla under various experimental conditions after 7 days   Sample Relative abundance (%) Day 0 0.12 Day 7 (Spiked) HET 0.21 HET+AOB+NOB 18 AOB+NOB 0.08 NOB 0.14 Day 7 (Control) HET 0.26 HET+AOB+NOB 10 AOB+NOB 0.08 NOB 0.13   a) b) 165   Known nitrifying bacteria identified through 16s RNA sequencing in 6:2 FTS biotransformation microcosms Phylum Class Order Family Genus Role in nitrification Reference Proteobacteria Betaproteobacteria Nitrosomonadales Nitrosomonadaceae Nitrosomonas AOB* Boch and Wagner, 2006 Proteobacteria Betaproteobacteria Nitrosomonadales Nitrosomonadaceae Nitrosovibrio AOB Boch and Wagner, 2006 Nitrospirae Nitrospira Nitrospirales Nitrospiraceae Nitrospira NOB and possibly AOB (Comamox) Daims et al., 2016 *AOB: ammonia-oxidizing bacteria; NOB: nitrite-oxidizing bacteria   Alpha diversity of microbial communities Richness Richness of a sample is the count of species, without taking into account the abundances of the species. However, a simple count of the number of species in a sample is usually an underestimate of the true number of species, because increasing the sampling effort (through counting more individuals, examining more sampling units, or sampling a larger area) increases the number of species observed (Gotelli and Chao, 2013). Therefore, asymptotic richness estimators are used for extrapolating species diversity to the (presumed) asymptote, beyond which additional sampling will not yield any new species. Most commonly used richness estimators are nonparametric (Colwell and Coddington, 1994), involving use of rare frequency counts to estimate the frequency of the missing species. Two of the most commonly used nonparametric richness estimators are: Chao1: Richness is estimated by the number of observed species (Sobs), added with a term that depends only on the observed number of singletons (a, species each represented by only a single individual) and doubletons (b, species each represented by exactly two individuals) (Colwell, 2009). 𝑆𝐶ℎ𝑎𝑜1 = 𝑆𝑜𝑏𝑠 + 𝑎22𝑏 166  Abundance-Based Coverage Estimators (ACE) richness: ACE estimator predicts the species richness based on the number of rarely occurring species. Usually, a frequency (k) of 10 or smaller is used as a cut-off for defining rate species.  𝑆𝐴𝐶𝐸 = 𝑆𝑎𝑏𝑢𝑛 +𝑠𝑟𝑎𝑟𝑒?̂?𝑟𝑎𝑟𝑒+𝑓1?̂?𝑟𝑎𝑟𝑒𝛾𝑟𝑎𝑟𝑒2    where, Srare is the number of rare species (the OTUs with ≤ 10 sequences) and Sabun is the number of abundant species (the OTUs with > 10 sequences). ?̂?𝑟𝑎𝑟𝑒 estimates the sample coverage and f1 is the number of species with single sequence. 𝛾𝑟𝑎𝑟𝑒2  is the square of the estimated coefficient of variation of the species relative abundances (Gotelli and Chao, 2013). Diversity Index Diversity indices are mathematical functions that considers evenness, in addition to richness and combine them in a single measure. The most commonly used diversity indices are the Shannon and Simpson diversity indices. Simpson index: This is a numerical index indicating the probability that that two unrelated strains sampled from the microbial communities will be placed into different classified groups (Hunter and Gaston, 1988) 𝐷 = 1 − ∑ 𝑝𝑖2𝑠𝑖=1 or preferably, 𝐷′ = (∑ 𝑝𝑖2𝑠𝑖=1)−1 where, species i constitutes pi proportions of the total individuals in a community of S species, Shannon index: This is a measure of the difficulty to predict the next individual strain in the sampled microbial communities.  𝐻 = − ∑ 𝑃𝑖 𝑙𝑛 𝜌𝑖𝑠𝑖=1 𝑜𝑟, 𝑝𝑟𝑒𝑓𝑒𝑟𝑎𝑏𝑙𝑦, 𝑒𝐻 where, species i constitutes pi proportions of the total individuals in a community of S species, 167  Both Shannon and Simpson diversities increase as richness increases, for a given pattern of evenness, and increase as evenness increases, for a given richness, their ranking of communities differ. Simpson diversity is less sensitive to richness and more sensitive to evenness than Shannon diversity, which, in turn, is more sensitive to evenness than richness (Colwell, 2009; Hill et al, 2003).  Comparing the Shannon diversity index of two samples using Hutcheson t-test (Gardener 2012). The Hutcheson t-test is a modified version of the classic t-test that can be used to compare diversity of two community samples using the Shannon diversity index (Hutcheson, 1970), that do not have replicates. The key formula that determines the variance of the Shannon index looks very similar to the classic t-test formula.  where H represents the Shannon diversity index for each of the two samples (subscripted a and b) and SH is the variance of each of the samples, calculated as follows,  Where, S is the species count, N is the total number of individuals (abundances), p is the proportion that each species makes towards the total. Assessing statistical significance  In order to determine the critical value of the t for a certain significance level, the degrees of freedom are calculated as follows, 168   where, N is the total abundance for each sample. The final value is close to the total abundance for the two samples added together. Following determination of t and df, critical value can be determined from a t-distribution table. If the calculated t-value exceeds the critical value, then the null hypothesis is rejected (meaning Shannon diversity differs statistically for the two samples). It is expected that most microbial communities will have large degrees of freedom. Therefore, the critical value for t will approach 2. The critical value for t at infinity is 1.96. The standard deviation of the Shannon index (square root of the variance) and multiplying by a factor of 2 can be defined as the confidence interval (Gardener, 2012).    169   Hutcheson t-test analysis for Shannon diversity index between 0 and 7 d samples under each treatment    day 7 day 7 Sample day 0 HET_spiked AOB+NOB_spiked NOB_spiked HET+AOB+NOB_spiked HET_control AOB+NOB_control NOB_control HET+AOB+NOB_control Total (abundance) 23289 16454 21826 26958 26307 23565 24233 20310 27287 Richness (Species count) 1870 1192 1360 1626 1320 1474 1353 1471 1383 H (Shannon Index) 4.94 4.58 4.41 4.59 4.10 4.57 4.22 4.65 4.39 S2H (variance) 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 CI (confidence interval) 0.029 0.032 0.029 0.026 0.027 0.028 0.028 0.029 0.025 Test conditions  Day 0 vs. HET spiked Day 0 vs. AOB+NOB spiked Day 0 vs. NOB spiked Day 0 vs. HET+AOB+NOB spiked Day 0 vs. HET control Day 0 vs. AOB+NOB control Day 0 vs. NOB control Day 0 vs. HET+AOB+NOB control t  16.58 25.91 18.19 42.28 18.51 35.75 14.09 29.24 df   36994.16 45089.72 48757.06 48764.33 46663.31 47346.91 43352.16 47686.89 Critical value  1.96 1.96 1.96 1.96 1.96 1.96 1.96 1.96 p  1.7347E-61 6.2009E-147 1.055E-73 2.06E-73 3.47956E-76 3.3843E-276 5.55306E-45 2.9153E-186 Test conditions  HET control vs. HET spiked AOB+NOB control vs. AOB+NOB spiked NOB control vs. NOB spiked HET+AOB+NOB control vs. HET+AOB+NOB spiked     t  0.54 9.47 3.34 15.52     df  35942.62 45732.02 44459.31 52879.30     Critical value  1.96 1.96 1.96 1.96     p  5.91E-01 2.88E-21 8.35E-04 3.42E-54     *The bold and italicized p values indicate null hypothesis is rejected (difference in Shannon diversity is statistically significant)   170   One-way ANOVA for comparing % 6:2 FTS remaining after 10 days under various treatment conditions          HET HET+AOB+NOB AOB+NOB NOB   Replicate 1 88.8 76.0 92.2 92.7   Replicate 2 98.4 85.0 83.4 95.0   Average 93.6 80.5 87.8 93.8          Anova: Single Factor             SUMMARY       Groups Count Sum Average Variance   HET 2 187.2126 93.60628 46.54864   HET+AOB+NOB 2 160.9778 80.48888 40.67599   AOB+NOB 2 175.6429 87.82144 38.41465   NOB 2 187.6636 93.83179 2.68083                 ANOVA       Source of Variation SS df MS F P-value F crit Between Groups 236.7526 3 78.91752 2.46002 0.20241 6.591382 Within Groups 128.3201 4 32.08003           Total 365.0727 7                  171   One-way ANOVA and Pairwise Comparison using Tucky test for total C4-C6 PFCA formation at day 10.  Sum of C4-C6 PFCAs at day 10 (nmol/L)     HET HET+AOB+NOB AOB+NOB NOB   Replicate 1 7.6 18.5 11.1 7.2   Replicate 2 7.3 26.9 7.5 6.1   Mean 7.5 22.7 9.3 6.7   Sample size (n) 2 2 2           Anova: Single Factor             SUMMARY       Groups Count Sum Average Variance   HET 2 14.9 7.5 0.0   HET+AOB+NOB 2 45.4 22.7 35.6   AOB+NOB 2 18.6 9.3 6.5   NOB 2 13.3 6.7 0.7                 ANOVA       Source of Variation SS df MS F P-value F crit Between Groups 339.3 3 113.1 10.5728541 0.022617 6.591382 Within Groups 42.8 4 10.7           Total 382.0761273 7          Tucky test for Pairwise Comparison            Tucky criterion                                                        For,       α = 0.05       𝑇 = 𝑞𝛼[𝑐(𝑛−𝑐)]√𝑀𝑆𝐸𝑛𝑖 qα[c(n-c)] = studentized range distribution based on c and (n-c) degrees of freedom           c = no. of treatments           n = total sample size MSE = Mean square error (from ANOVA table) ni = sample size of treatment group 172  qα[c(n-c)] =   5.757       ni = 2       MSE = 10.6969              T = 13.314              Pair Absolute difference between means Criterion (T) significant at 0.05    HET to AOB+NOB 1.9 13.314 No    AOB+NOB to NOB 2.7 13.314 No    NOB to HET+AOB+NOB 16.0 13.314 Yes    HET to NOB 0.8 13.314 No    HET to HET+AOB+NOB 15.2 13.314 Yes    AOB+NOB to HET+AOB+NOB 2.7 13.314 No              173   Comparison of experimental conditions, and major outcomes of previous and current 6:2 FTS biotransformation studies under aerobic conditions  a: The coefficient of determination (R2) varied between 0.8 – 0.5 for 7 data points. bN/A: not available  Degradation media  Duration (day) Spiking concentration Stable products (yield in mol%) Half-life (day) Reference Landfill leachate and sediment 60 ~1.5 µmol/L PFBA (1 – 2%), PFPeA (5 – 7%), PFHxA (5%) 90 – 110a  Chapter 4 of this study Inoculum from landfill leachate ditch 10 ~1.7 µmol/L PFBA (0.3%), PFPeA (0.1%), PFHxA (0.7%) N/Ab Chapter 5 of this study Landfill leachate and sediment 90 ~3 µmol/L PFBA (0.6%), PFPeA (5.6%), PFHxA (3.1%), 5:3 FTCA (0.2%), 5:2 sFTOH (0.6 %) ~86  (R2 = 0.86) Chapter 3 of this study Activated sludge 90 2.2 mg/L 5:3 FTCA (0.12%), PFBA (0.14%), PFPeA (1.5%), and PFHxA (1.1%)  ~2 years Wang et al., 2011 Gordonia sp. strain NB4-1Y 5  N/A 5:3 FTCA, 6:2 FTCA, 6:2 FTUCA and 5:3 FTUCA N/A Van Hamme et al., 2013 Activated sludge Up to 40 weeks 500 mg/L No observable biotransformation - Ochoa-Herrera et al., 2016 River sediment 90 2.8 nmol/L 5:3 FTCA (16%), PFPeA (21%) and PFHxA (20%) <5  Zhang et al., 2017 Gordonia sp. strain NB4-1Y  7 N/A 5:2 FT ketone (43.9%), 5:2 sFTOH (8.97%), and 6:2 FTOH (4.14 %) <7 Shaw et al., 2019 174  Appendix E  Supplemental Information on Photolysis of 6:2 Fluorotelomer Sulfonate (6:2 FTS) in Landfill Leachate Under Simulated Sunlight (Chapter 6)    C C C C C C C CFFFF FFFF FF FFFSOOOHHHHH   Figure E.1 Chemical structure of 1H, 1H, 2H, 2H-perfluorooctane sulfonate (6:2 FTS) showing the perfluoroalkyl tail (green), ethyl moiety with abstractable H atoms (red) and sulfonate group (magenta) (ACD/ChemSketch™ Freeware)            175   Standards of per- and polyfluoroalkyl substances (PFASs) and their suppliers Analyte Acronym Chemical Formula CAS Source Perfluoro-n-butanoic acid PFBA C3F7COOH 377-22-4 Wellington Laboratories, Canada Perfluoro-n-pentanoic acid PFPeA C4F9COOH 2706-90-3 Wellington Laboratories, Canada Perfluoro-n-hexanoic acid PFHxA C5F11COOH 307-24-4 Wellington Laboratories, Canada Perfluoro-n-heptanoic acid PFHpA C6F13COOH 375-85-9 Wellington Laboratories, Canada Perfluoro-n-octanoic acid PFOA C7F15COOH 335-67-1 Wellington Laboratories, Canada 3-Perfluoropentyl propanoic acid 5:3 FTCA C8H5F11O2 914637-49-3 Wellington Laboratories, Canada Sodium 1H,1H,2H,2H-perfluorooctane sulfonate (6:2) 6:2 FTS C8H4F13SO3Na 27619-97-2 Synquest Laboratories, Florida, USA, Wellington Laboratories, Canada  Sodium 1H,1H,2H,2H-perfluoro-1-[1,2-13C2] octane sulfonate (6:2) M2_6:2 FTS - - Wellington Laboratories, Canada Perfluoro-n-[2,3,4-13C ] butanoic acid M3PFBA - - Wellington Laboratories, Canada Perfluoro-n-[1-13C5] hexanoic acid MPFHxA - - Wellington Laboratories, Canada Perfluoro-n-[13C8] octanoic acid M8PFOA - - Wellington Laboratories, Canada     Chemicals used for phosphate buffer and stock solution preparation and their suppliers Chemical  Supplier Purity Potassium phosphate, Monobasic (KH2PO4) Millipore-Sigma, Canada ≥ 99% Dipotassium hydrogen orthophosphate (K2HPO4) VWR BDH Chemicals, Canada ≥ 98% Humic Acid Sigma-Aldrich, USA <90% Sodium Nitrate (NaNO3) Fisher Scientific, USA ≥ 99%  176    Figure E.2 Spectrogram of the 1000 W metal halide lamp used for irradiation experiments obtained from the lamp manufacturer.   Figure E.3 a) Soaking chamber with b) metal halide lamp and sample tray. 177    Figure E.4 Temperature profile inside the soaking chamber for a typical 72-hour irradiation experiment.              010203040500 10 20 30 40 50 60 70 80Temperature (°C)Time (Hours)178   Instrumental method parameters for analysis of PFCAs and 6:2 FTS by LC-MS/MS  Instruments Agilent Technologies 1200 liquid chromatography 6410 Triple Quad mass spectrometer. Operated in the negative ion multiple reaction monitoring mode. Analytical Column Waters Xterra C18MS analytical column, 100 mm length, 2.1 mm ID, 3.5 μm particle size;  Guard Column Waters Xterra C18MS analytical column, 30 mm length, 2.1 mm ID, 3.5 μm particle size LC conditions Column temperature (oC): 50 Maximum Pressure (bar): 345 Flow rate (mL/min): 0.7 Mobile Phases A: 10 mM Ammonium acetate and 10 mM acetic acid in water B: Acetonitrile Solvent gradient Profile Time (min) A (%) B (%) 0 80 20 6 40 60 7 10 90 8 10 90 9 80 20 12 80 20  MS conditions Source temperature (oC): 120 Desolvation temperature (oC): 325 Capillary voltage (kV): 3.5 Injection Volume: 4 μL Monitored Ion Transitions Analytes Fragmentor voltage (V) Collision energy (eV) Parent to product ion transition PFBA 28 0 213.0 -> 169.0 PFPeA 20 0 263.0 -> 219.0 PFHxA 30 0 313.0 -> 269.0 PFHpA 30 0 363.0 -> 319.0 PFOA 56 2 413.0 -> 369.0 5_3 FTCA 70 2 341.0 -> 237.0 6_2 FTS 130 2 427.0 -> 407.0 M3FPBA 28 0 216.0-> 172.0 M2_6:2 FTS 130 2 429.0 -> 81.0 MPFHxA 30 0 315.0 -> 270.0 M8PFOA 56 2 413.0 -> 369.0  Calibrations  6:2 FTS was quantified using a 6-point calibration in the working concentration range (2 to 250 µg/L) with M2_6:2 FTS as an internal standard. All the other compounds were quantified using a 6-point calibration in the working concentration range (0.1 to 20 µg/L) with M3PFBA, MPFHxA and M8PFOA as injection internal standards. The limits of detection (LODs) (E4) were calculated using based on signal-to-noise ratios of 3 determined at the lowest calibration limit.    179   The instrumental detection limit (µg/L) of target analytes in deionized water Analytes Internal Standard Limit of detection (LOD)a PFBA M3PFBA 0.5 PFPeA M3PFBA 0.2 PFHxA MPFHxA 0.3 PFHpA MPFHxA 0.2 PFOA M8PFOA 0.1 5:3 FTCA MPFHxA 0.1 6:2 FTS M2_ 6:2 FTS 0.5 asignal to noise ratio of 3 determined at lowest calibration level     Figure E.5 Plot showing natural logarithm of 6:2 FTS concentration in leachate with time, used for pseudo-first order reaction kinetics calculation.  y = -0.0039x - 0.0542R² = 0.9349-0.35-0.3-0.25-0.2-0.15-0.1-0.0500 10 20 30 40 50 60 70 80Ln (concentration)Time (hours)

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