ASSESSMENT OF WATER TREATMENT TECHNOLOGIES FOR PER-AND POLYFLUOROALKYL SUBSTANCES (PFAS) IN MULTIPLE MATRICES By Vanessa Maldonado A DISSERTATION Submitted to Michigan State University in partial fulfillment of the requirements for the degree of Chemical Engineering – Doctor of Philosophy 2022 ABSTRACT ASSESSMENT OF WATER TREATMENT TECHNOLOGIES FOR PER-AND POLYFLUOROALKYL SUBSTANCES (PFAS) IN MULTIPLE MATRICES By Vanessa Maldonado The ubiquitous presence of per-and polyfluoroalkyl substances (PFAS) in the environment resulted in extensive water contamination that poses a significant risk to human health and biota. Continu- ous research efforts aim to develop efficient treatment technologies to treat PFAS in water, break the PFAS accumulation cycle in the environment, and improve the efficiency of emerging tech- nologies. In this thesis work, selected treatment technologies including electrochemical oxidation and dielectrophoresis-enhanced adsorption were used to assess and advance the state-of-the-art for PFAS remediation in multiple matrices, not previously addressed. A boron-doped diamond (BDD) flow-through cell was used to evaluate the electrochemical oxi- dation of perfluoroalkyl acids (PFAAs) in landfill leachates. Multiple leachates with a concentration of individual PFAAs in the range of 10 2 – 10 4 ng/L were treated. The effect of current density and variability of the composition of leachates was investigated. Non-detect levels and >90% removal of perfluorooctane sulfonate (PFOS) and perfluorooctanoic acid (PFOA) were reached for all leachates treated, respectively. Although high removal efficiencies for long-chain PFAAs were obtained, high concentrations of short-chain PFAAs were generated and associated with the transformation of perfluoroalkyl acid (PFAA) precursor compounds. In the second part of this thesis research, the oxidative transformation of PFAA-precursors, typically present in leachates, was addressed for the first time. Target and suspect PFAS were identified in a landfill leachate and their concentrations during the electrochemical treatment were quantified over time. Liquid chromatography quadrupole time-of-flight mass spectrometry (LC- QToF) measurements of the leachate identified 53 PFAS compounds and 19 PFAS classes. Multiple PFAS were reported for the first time in landfill leachates. The evaluation of the intermediate and final products generated during the electrochemical treatment showed evidence of known electrochemical degradation pathways. Coupling destructive technologies (e.g., electrochemical oxidation) with concentration tech- nologies (e.g., ion exchange (IX), adsorption) in a treatment train approach could reduce the treatment cost of destructive technologies and increase their feasibility. Therefore, in the next part of this work, electrochemical oxidation of PFAAs from the concentrated waste of IX still bottoms was assessed at laboratory and semi-pilot scales. The concentrated waste resulted from the treat- ment of PFAAs-impacted groundwater with IX resins. Multiple current densities were evaluated at the laboratory scale and the optimum current density was used at the semi-pilot scale. The results at the laboratory and semi-pilot scales allowed for >99% and >94% removal of total PFAAs with 50 mA/cm2 , respectively. Defluorination values, energy consumption, and implications were discussed. The third matrix addressed for PFAS remediation was drinking water. Dielectrophoresis- enhanced adsorption was used for the removal of low concentrations of PFOA. This study introduced a coaxial-electrode cell (CEC) that allowed for the generation of a non-uniform electric field to enhance the adsorption of PFOA. Experiments were performed in batch and continuous-flow modes. The dielectrophoresis-enhanced adsorption in batch mode resulted in a 4, 7, and 8-fold increase in the removal of PFOA with 5, 25, and 50 V, respectively, when compared to adsorption only. The performance of the CEC in continuous-flow mode allowed for an increase of up to 2.4-fold in the PFOA removal with 25 V. The results highlighted the benefits of using a dielectrophoresis-enhanced adsorption process for the removal of PFOA from water. Overall, results from this thesis contribute to the understanding of the electrochemical degrada- tion of PFAS in multiple matrices and introduce an alternative process to enhance the widely used adsorption technology for PFAS removal. Treatment implications of each matrix are discussed and provide a clear baseline for future research, development, and scale-up of treatment technologies for PFAS remediation. Copyright by VANESSA MALDONADO 2022 To Chris. For his kindness, love, and patience. v ACKNOWLEDGEMENTS I would like to express my deepest gratitude to my advisor Dr. Qi Hua Fan for his guidance and support. His motivation, encouragement, and trust allowed me to conclude multiple projects that shaped my skills, knowledge, and abilities to become a scientist. The work presented in this dissertation would have not been possible without the guidance and input of several people. I am thankful to my committee members, Dr. Thomas Schuelke, Dr. Scott Barton, Dr. Greg Swain, Dr. Richard Lunt, and Dr. Volodomyr Tarabara for their scientific advice, guidance, and support. I am infinitely grateful to Dr. Jennifer Field from the Department of Chemistry at Oregon State University for her extensive input and time. Collaborating with one of the pioneers and experts in PFAS research was an invaluable learning experience. It was the greatest honor to work with her and certainly, the greatest memory of my path as a Ph.D. student. I also want to recognize our collaborators and friends, Dr. Sibel Uludag-Demirer (Biosystems & Agricultural Engineering), Dr. Anthony Schilmiller (Mass Spectrometry and Metabolomics Core), and Daniel Holmes (Chemistry Department). I would like to give them special thanks for the training and support provided for samples processing and analysis. I am grateful to my undergraduate students Greg Landis, Emma Davis, Kristina Harbin, Theresa Waeltermann, and Callaghan Tyson Mayer for their valuable contributions to various projects. I also want to thank Mary Ensch and Lexi Rogien, former graduate student members of the Fraunhofer Center. Their company certainly alleviated the loneliness of graduate school and left beautiful memories to remember. I will also like to express my gratitude to Michael Becker, Dr. Suzanne Witt, and Dr. Cory Rusinek for their collaboration and support in the various Fraunhofer projects. Special thanks to Michael for opening the doors of the center to this server of the world. I am truly thankful to my love Chris for the infinite number of hours dedicated to actively listening to my graduate school stories, anecdotes, challenges, and for his unconditional love and support. Thanks for being my fan number one and the best events coordinator/coffee maker/organizer. I want to honor my family in Ecuador, especially my mom and dad, for their love and encouragement vi regardless of the distance. Additional thanks to my friends in Ecuador for staying present in my life and for their support during these years. Finally, I would like to thank the City of Grand Rapids (Michigan, U.S.), the Fraunhofer Center Midwest, Division for Coatings and Diamond Technologies, and the Environmental Research and Education Foundation (EREF) for their financial support. vii TABLE OF CONTENTS LIST OF TABLES . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . xi LIST OF FIGURES . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . xiii CHAPTER 1 INTRODUCTION . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1 1.1 PFAS background and classification . . . . . . . . . . . . . . . . . . . . . . . . . 2 1.2 Physical and Chemical Properties . . . . . . . . . . . . . . . . . . . . . . . . . . . 5 1.3 Synthesis of PFAS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5 1.3.1 Electrochemical fluorination . . . . . . . . . . . . . . . . . . . . . . . . . 5 1.3.2 Telomerization . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6 1.4 Products and Uses . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6 1.5 Environmental Persistence and Toxicity . . . . . . . . . . . . . . . . . . . . . . . 7 1.6 Occurrence of PFAS in water and wastewater . . . . . . . . . . . . . . . . . . . . 8 1.6.1 Wastewater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8 1.6.2 Surface water . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9 1.6.3 Drinking water . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9 1.6.4 Groundwater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10 1.7 Environmental Regulations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10 1.8 Switching to green alternatives . . . . . . . . . . . . . . . . . . . . . . . . . . . . 11 1.9 The cyclical problem of PFAS disposal . . . . . . . . . . . . . . . . . . . . . . . . 11 1.10 Treatment technologies for PFAS . . . . . . . . . . . . . . . . . . . . . . . . . . . 12 1.11 Destructive technologies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13 1.11.1 Electrochemical oxidation . . . . . . . . . . . . . . . . . . . . . . . . . . 13 1.11.1.1 Boron-doped diamond . . . . . . . . . . . . . . . . . . . . . . . 14 1.11.1.2 Electrochemical degradation mechanisms for PFAS . . . . . . . . 15 1.11.2 Ozonation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16 1.11.3 Activated Persulfate . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16 1.11.4 Plasma treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17 1.11.5 Incineration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17 1.12 Non-destructive treatment technologies . . . . . . . . . . . . . . . . . . . . . . . . 18 1.12.1 Adsorption . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18 1.12.2 Ion exchange . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 19 1.12.3 Foam fractionation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20 1.13 Research motivation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20 1.14 Research Objectives . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 21 1.14.1 General Objective . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 21 1.14.2 Specific Objectives . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 21 BIBLIOGRAPHY . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22 viii CHAPTER 2 A FLOW-THROUGH CELL FOR THE ELECTROCHEMICAL OXIDATION OF PERFLUOROALKYL SUBSTANCES IN LANDFILL LEACHATES . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 31 2.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 32 2.2 Materials and methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 34 2.2.1 Materials . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 34 2.2.2 Landfill leachates . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 34 2.2.3 Electrochemical oxidation setup . . . . . . . . . . . . . . . . . . . . . . . 36 2.2.4 Analytical methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 37 2.3 Results and discussion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 39 2.3.1 Performance of the BDD flow-through cell . . . . . . . . . . . . . . . . . . 39 2.3.2 Influence of current density on the electrochemical treatment of PFAAs in landfill leachates . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 43 2.3.3 Electrochemical treatment of various leachates . . . . . . . . . . . . . . . 50 2.3.4 Perchlorate generation in leachates . . . . . . . . . . . . . . . . . . . . . . 54 2.4 Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 56 BIBLIOGRAPHY . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 58 CHAPTER 3 ELECTROCHEMICAL TRANSFORMATIONS OF PERFLUOROALKYL ACID (PFAA) PRECURSORS AND PFAAS IN LANDFILL LEACHATES . . . . . . . . . . . . . . . . . . . . . . . . . . . 63 3.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 64 3.2 Experimental section . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 67 3.2.1 Chemicals and reagents . . . . . . . . . . . . . . . . . . . . . . . . . . . . 67 3.2.2 Sample collection . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 68 3.2.3 Electrochemical oxidation setup . . . . . . . . . . . . . . . . . . . . . . . 70 3.2.4 Analytical methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 71 3.2.4.1 PFAS quantification . . . . . . . . . . . . . . . . . . . . . . . . 72 3.3 Results and discussion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 77 3.3.1 PFAS characterization in untreated L1 . . . . . . . . . . . . . . . . . . . . 77 3.3.1.1 PFAS detected with LC-QToF . . . . . . . . . . . . . . . . . . . 77 3.3.1.2 PFAS contribution from TOP Assay . . . . . . . . . . . . . . . . 83 3.3.2 PFAS transformations during electrochemical oxidation of L1 . . . . . . . 84 3.3.3 Electrochemical degradation pathways of L1 . . . . . . . . . . . . . . . . . 94 3.3.4 Fluorine mass balance . . . . . . . . . . . . . . . . . . . . . . . . . . . . 98 3.3.5 Energy consumption and total organic carbon removal . . . . . . . . . . . 99 3.3.6 Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 100 APPENDICES . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 102 APPENDIX 3A CHARACTERIZATION OF BDD ANODES . . . . . . . . . . . . 103 APPENDIX 3B PFAS CHARACTERIZATION OF L1 . . . . . . . . . . . . . . . . 106 BIBLIOGRAPHY . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 112 CHAPTER 4 LABORATORY AND SEMI-PILOT SCALE STUDY ON THE ELECTROCHEMICAL TREATMENT OF PERFLUOROALKYL ACIDS FROM ION EXCHANGE STILL BOTTOMS . . . . . . . . . . . . . . . . 119 ix 4.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 120 4.2 Materials and Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 121 4.2.1 Materials . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 121 4.2.2 Electrochemical Oxidation Setup . . . . . . . . . . . . . . . . . . . . . . . 121 4.2.3 Electrochemical Experiments . . . . . . . . . . . . . . . . . . . . . . . . . 124 4.2.4 Analytical Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 124 4.3 Results and Discussion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 126 4.3.1 Laboratory Scale Evaluation . . . . . . . . . . . . . . . . . . . . . . . . . 126 4.3.1.1 General Observations . . . . . . . . . . . . . . . . . . . . . . . 126 4.3.1.2 Influence of Current Density on PFAAs Removal . . . . . . . . . 127 4.3.1.3 Electrochemical Treatment of Real Still Bottoms . . . . . . . . . 131 4.3.2 Semi-Pilot-Scale Evaluation . . . . . . . . . . . . . . . . . . . . . . . . . 133 4.3.3 Perchlorate Formation during Electrochemical Treatment . . . . . . . . . . 137 4.3.4 Treatment Efficiency and Energy Consumption . . . . . . . . . . . . . . . 138 4.4 Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 140 BIBLIOGRAPHY . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 142 CHAPTER 5 DIELECTROPHORESIS-ENHANCED ADSORPTION FOR THE REMOVAL OF PFOA FROM WATER . . . . . . . . . . . . . . . . . . . . 147 5.1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 148 5.2 Materials and methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 149 5.2.1 Materials . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 149 5.2.2 Fabrication of carbon-coated electrodes . . . . . . . . . . . . . . . . . . . 150 5.2.3 Characterization of electrodeposited electrodes . . . . . . . . . . . . . . . 150 5.2.4 Theory of dielectrophoresis drift of dipole . . . . . . . . . . . . . . . . . . 151 5.2.5 Dielectrophoresis-enhanced adsorption cell . . . . . . . . . . . . . . . . . 153 5.2.6 Analytical methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 154 5.3 Results and discussion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 154 5.3.1 Characterization of electrodes . . . . . . . . . . . . . . . . . . . . . . . . 154 5.3.2 Mechanisms of dielectrophoresis enhanced PFAS adsorption . . . . . . . . 155 5.3.3 Dielectrophoresis effect on the adsorption of PFOA in batch mode . . . . . 157 5.3.4 Continuous flow-operation of the CEC . . . . . . . . . . . . . . . . . . . . 159 5.4 Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 161 BIBLIOGRAPHY . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 162 CHAPTER 6 CONCLUSIONS AND FUTURE DIRECTIONS . . . . . . . . . . . . . . . 167 6.1 Summary and Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 168 6.2 Challenges encountered . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 170 6.3 Future Directions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 171 x LIST OF TABLES Table 1.1. Short-chain and long-chain PFCAs and PFSAsa . . . . . . . . . . . . . . . . . . 4 Table 2.1. Characterization of leachate samples . . . . . . . . . . . . . . . . . . . . . . . . 34 Table 2.2. Initial concentrations of PFAS, quantified in different leachate samples. All the values are shown in ng/L . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 35 Table 2.3. Gradient solvent program for the HPLC . . . . . . . . . . . . . . . . . . . . . . 38 Table 2.4. Calibration standards used for PFAS detection . . . . . . . . . . . . . . . . . . . 39 Table 2.5. Surrogate used for PFAS detection . . . . . . . . . . . . . . . . . . . . . . . . . 40 Table 2.6. Comparison of current normalized rate constants and energy per order values for the electrochemical oxidation of PFOA and PFOS among various studies. All studies were performed in batch mode. All studies used a parallel-plate cell configuration, except for a in this work that used a flow-through cell . . . . . . . 42 Table 2.7. Values of kinetic rate constants for COD evolution during the electrochemical oxidation of leachate L1 with multiple current densities . . . . . . . . . . . . . . 50 Table 2.8. Initial concentration of PFCAs, PFSAs, and total PFAS of the leachates treated in this study . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 51 Table 2.9. Values of zero order kinetic rate constants for perchlorate generation during the electrochemical oxidation of leachate L1 with multiple current densities . . . . . 55 Table 3.1. Surrogate standards used for target PFAS analysis . . . . . . . . . . . . . . . . . 66 Table 3.2. Target PFAS, acronym, and surrogate standards for analysis by LC-QToF. . . . . 67 Table 3.3. Suspect PFAS detected in leachate L1. . . . . . . . . . . . . . . . . . . . . . . . 69 Table 3.4. Target PFAS, acronym, accuracy (% recovery), precision (% RSD), and limits of detection and quantification in landfill leachate by LC-QToF. The ‘*’ indicates that the surrogate standard was used in place of the target to estimate the LOD/LOQ. b ND indicates no surrogate available and target PFAS was in background leachate. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 74 Table 3.5. Average suspect PFAS concentrations ± standard error found in untreated L1. . . 79 xi Table 3.6. Characterization for leachate L1a . . . . . . . . . . . . . . . . . . . . . . . . . . 80 Table 3.7. Influent PFAS concentrations in L1 (ng/L and nM) ± standard error and summed masses ± propagated standard error . . . . . . . . . . . . . . . . . . . 83 Table 3.8. Percentage removal of PFAAs in L1 after electrochemical treatment with multiple current densities. Negative values represent increase in concentration . . 100 Table 3B.1. PFAS characterization of L1. Analytes include target and suspect PFAS. "n" represents the number of C with at least 1 F. Concentration values correspond to the average ± standard error. . . . . . . . . . . . . . . . . . . . . . . . . . . . . 107 Table 4.1. Characterization of the synthetic still bottoms solution used for the electrochemical treatment of PFAAs in both laboratory and semi-pilot scales . . 122 Table 4.2. Characterization of the real still bottoms solution used for the electrochemical treatment of PFAAs in laboratory scale . . . . . . . . . . . . . . . . . . . . . . 122 Table 4.3. Specifications of the electrochemical setup at laboratory scale and semi-pilot scale123 Table 4.4. Calibration standards used for PFAS detection . . . . . . . . . . . . . . . . . . . 125 Table 4.5. Values of fluoride pseudo-first order generation rate constants during the electrochemical treatment of PFAS in still bottoms . . . . . . . . . . . . . . . . 127 Table 4.6. Values of surface area normalized pseudo-first order degradation rate constants for the electrochemical treatment of PFAS in from a synthetic still bottoms solution128 xii LIST OF FIGURES Figure 1.1. Main PFAS classification. Adapted from [1] . . . . . . . . . . . . . . . . . . . 3 Figure 1.2. Example of transformation of PFAA-precursors to PFAAs. Adapted from [1]. . 4 Figure 1.3. Chemical structure of (a) PFOA and (b) PFOS . . . . . . . . . . . . . . . . . . 4 Figure 1.4. Influent and effluent concentrations (ng/L) of selected PFAS compunds in PFCA and PFSA groups in WWTPs. "I" stands for influent. "E" stands for effluent. Reprinted with permission from [15] . . . . . . . . . . . . . . . . . . 8 Figure 1.5. Concentration of PFAS in drinking water (ng/L). Reprinted with permission from [15] . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9 Figure 1.6. Pollution cycle of PFAS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12 Figure 1.7. Electrochemical oxidation cell . . . . . . . . . . . . . . . . . . . . . . . . . . 13 Figure 1.8. Oxidation mechanisms in the electrochemical oxidation process . . . . . . . . . 14 Figure 1.9. Mechanism of competitive adsorption of long chain, short chain PFAS and organic matter (OM). Reprinted with permission from [15] . . . . . . . . . . . 18 Figure 2.1. Schematic of BDD flow-through cell . . . . . . . . . . . . . . . . . . . . . . . 36 Figure 2.2. Experimental set up for the electrochemical oxidation of PFAAs with a BDD flow-through cell. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 37 Figure 2.3. (a) Decrease in concentration of PFOA and PFOS, and (b) byproducts of PFOA and PFOS oxidation over time during the electrochemical treatment of synthetic solutions with a BDD flow-through cell. Individual initial concentrations of PFOA and PFOS were 70 μg/L. Current density applied = 50 mA/cm2 . Solutions were prepared with 10 mM Na2 SO4 . . . . . . . . . . . . 41 Figure 2.4. Schematic of BDD parallel-plate cell . . . . . . . . . . . . . . . . . . . . . . . 41 Figure 2.5. Concentration of total PFAAs over time during the electrochemical oxidation of leachate L1 with a BDD flow-through cell. The applied current densities were: (a) 50 mA/cm2 , (b) 100 mA/cm2 , (c) 150 mA/cm2 , and (d) 200 mA/cm2 . Samples were spiked with [PFOA]0 ≈ 25 μg/L and [PFOS]0 ≈ 15 μg/L. . . . . 44 xiii Figure 2.6. Concentration of (a) PFSAs and (b) PFCAs during the electrochemical oxidation of leachate L1 with multiple current densities. . . . . . . . . . . . . . 45 Figure 2.7. Fraction of molar F relative to 𝑡 = 0 in PFCAs during the electrochemical oxidation of leachate L1 with (a) 50 mA/cm2 , (b) 100 mA/cm2 , (c) 150 mA/cm2 , and (d) 200 mA/cm2 . . . . . . . . . . . . . . . . . . . . . . . . . 47 Figure 2.8. Fraction of molar F relative to 𝑡 = 0 in PFSAs during the electrochemical oxidation of leachate L1 with (a) 50 mA/cm2 , (b) 100 mA/cm2 , (c) 150 mA/cm2 , and (d) 200 mA/cm2 . . . . . . . . . . . . . . . . . . . . . . . . . 48 Figure 2.9. Electrochemical degradation of PFBA (1 mg/L) with a current density of 150 mA/cm2 using Na2 SO4 and NaCl as Chapter2porting electrolytes. . . . . . . . . 49 Figure 2.10. Evolution of concentration of COD with respect to t=0 over time during the electrochemical oxidation of leachate L1 with multiple current densities. [PFOA]0 ≈ 28 μg L – 1 ; [PFOS]0 ≈ 18 μg L – 1 . . . . . . . . . . . . . . 49 Figure 2.11. Box and whisker plot for: (a) initial concentrations of PFAAs detected in five different leachate samples (L2−L6), and (b) removal efficiency (%) of leachates L2−L6 after 2 h of electrochemical oxidation with an applied current density of 150 mA/cm2 . Removal efficiency is between +100 and −100% . Ends of the boxes represent the first and third quartiles, horizontal line inside the box represent the median, whiskers represent minimum and maximum values. Samples were not spiked. . . . . . . . . . . . . . . . . . . . . . . . . . 51 Figure 2.12. Concentration of total PFAAs over time during the electrochemical oxidation of multiple landfill leachates with a BDD flow-through cell. The landfills correspond to: (a) L2, (b) L3, (c) L4, (d) L5, and (e) L6. The applied current density was 150 mA/cm2 . Samples were not spiked. . . . . . . . . . . . . . . . 53 Figure 2.13. Concentration of perchlorate over time during the electrochemical oxidation of landfill leachates with multiple current densities. [PFOA]0 ≈ 28 μg/L; [PFOS]0 ≈ 18 μg/L. Samples correspond to leachate L1. . . . . . . . . . . . . 55 Figure 3.1. Electrochemical oxidation setup used for the electrochemical treatment of leachate L1 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 70 Figure 3.2. Process diagram of the extraction method used for leachates . . . . . . . . . . . 71 Figure 3.3. Average target PFAS concentrations ± standard error in untreated leachate L1 based on measurement of n = 3 replicates measured by LC-QToF. Colors represent different PFAS classes. Only PFAS with concentrations >LOQ are represented. PFAS with concentrations 2.74 V/SHE)[57]. According to Comninellis et al., during the water electrolysis, BDD anodes promote the production of adsorbed hydroxyl radicals •OH (Eq. 1.1), which reacts with the organic molecules Figure 1.8. Oxidation mechanisms in the electrochemical oxidation process 14 (Eq. 1.2) [45]. At the same time, the oxygen evolution reaction (Eq. 1.3) takes place as a secondary competing reaction of the degradation process. BDD + H2 O → BDD( •OH) + H+ + e− (1.1) BDD ( •OH) + R → BDD + CO2 + H2 O (1.2) 1 BDD ( •OH) → (BDD) + O2 + H+ + e− (1.3) 2 1.11.1.2 Electrochemical degradation mechanisms for PFAS The electrochemical oxidation mechanism of PFCAs and PFSAs has been widely documented [15, 36, 58, 59]. For PFCAs, two possible electrochemical pathways have been reported: 1) H/F exchange on the C−F bonding and 2) the unzipping mechanism. The latter undergoes a stepwise elimination of CF2 moieties after an initial electron transfer of an electron from the head group of the PFAA molecule to the anode (Eq. 1.4) and includes: decarboxylation (Eq. 1.5), hydroxylation (Eq. 1.6), elimination (Eq. 1.7), and hydrolysis (Eq. 1.8) that breaks PFAAs in smaller fractions and releases CF2 moeities [60–63]. C𝑛 F2𝑛+1 COO− → C𝑛 F2𝑛+1 COO• + e− (1.4) C𝑛 F2𝑛+1 COO• → C𝑛 F2𝑛+1 • + CO2 + e− (1.5) C𝑛 F2𝑛+1 • + H2 O → C𝑛 F2𝑛+1 OH + H• (1.6) C𝑛 F2𝑛+1 OH → C𝑛−1 F2𝑛−1 COF + F− + H+ (1.7) C𝑛−1 F2𝑛−1 COF + H2 O → C𝑛−1 F2𝑛−1 COO− + F− + 2 H+ (1.8) The degradation of PFSAs has been reported to occur through desulfonation (Eq. 1.9), H/F exchange (Eq. 1.10), and chain shortening (Eq. 1.6, 1.7, 1.8). The H/F exchange can lead to 15 poly-fluorinated compounds and the chain shortening can originate multiple short-chain PFCAs. [15, 58, 59] C𝑛 F2𝑛+1 SO3 2− → C𝑛 F2𝑛+1 − + SO3 •− (1.9) C𝑛 F2𝑛+1 − + H2 O+ → C𝑛 F2𝑛+1 OH + H• (1.10) 1.11.2 Ozonation Ozonation treatment is based on the addition of ozone through diffusers to the solution to be treated. The strong oxidative properties of ozone allow for the oxidation of species through two different pathways: Direct attack of molecular ozone (Eq. 1.11) or generation of hydroxyl radicals upon decomposition of ozone (Eq. 1.12)[64]. O3 + OH− → − HO2 − + H+ (1.11) O3 + HO2 − → − HO• + O2 •− + O2 (1.12) A limited number of studies have been conducted for the degradation of PFAS, which has shown to occur only under alkaline conditions (eg. pH 11) [1, 64]. Additional oxidants, such as persulfate and iron-oxide based catalysts were used to increase the efficiency of removal [1, 65]. Ozone and UV treatment or air fractionation were combined to enhance the degradation of PFAS [66]. Ozonation processes can generate potential toxic transformation byproducts as the organic compounds are not completely oxidized to carbon dioxide and water. Therefore, to avoid non-desired intermediates, the technology is usually coupled with UV oxidation. 1.11.3 Activated Persulfate Oxidation of PFAS with activated persulfate (S2 O8 ) has become of interest due to its high oxidation potential (E0 = 2.1 V). In addition, with the influence of UV light, temperature, microwave energy, alkaline pH or hydrogen peroxide, S2 O8 generates sulfate radicals (SO4 • ) which also act as strong 16 oxidants of PFAS [67–69]. Persulfate oxidation has been applied to the degradation of PFAS in synthetic solutions [70], groundwater [69] and AFFF solutions [71]. Although the oxidation with heat-activated persulfate has been successful in treating PFCAs, it has shown limited or no degradation for PFSAs [68, 71]. Additionally, the transformation of persulfate to sulfate radicals leads to the release of H+ [71], which drastically drops the pH of the solution ( 1.5). Consequently, the treated solution should be restored to neutral pH, reducing the feasibility of the treatment. 1.11.4 Plasma treatment Plasma-based water treatment uses an electrical discharge to convert water into highly reactive species including •OH, O, H• , HO2 • , O2 • , H2 , O2 , H2 O2 , through energetic electrons in the plasma – (eaq ) [72]. The electrical discharge can be generated between two electrodes, one high voltage located above the liquid interface and another grounded which is in contact with the water to treat [73]. Through the generation of bubbles with diffusers, surface-active PFAS are driven to the water-air interface in the form of foam. The previous step allows PFAS to be directly exposed to the plasma at the interface which generates highly oxidative and reductive species that allow for a fast degradation of PFAS [72, 74]. The degradation of PFAS with plasma has been evaluated in multiple matrices, including synthetic solutions, groundwater, and wastewater [72, 75–77]. 1.11.5 Incineration Incineration is a chemical technology based on the combustion of materials/substances at high temperatures. An estimate of 12% or 34 million tons, of the municipal solid waste in the US is incinerated annually [78]. In addition to municipal waste incineration, there are hundreds of facilities for sewage sludge, hazardous waste, and medical waste incineration [79, 80]. The incineration of PFAS-based materials or waste containing PFAS has been applied in the US [42, 81]. However, the fate and transport of PFAS during incineration are not yet well understood [34, 82]. In addition, complete combustion of PFAS requires temperatures of at least 1000 °C [83]. Although, previous studies found that specific PFAS, such as PFOA and PFOS can be broken down with 17 incineration [84, 85], the full scope of potential PFAS byproducts that could form during the combustion of PFAS yet to be addressed. Public concerns have raised regarding the potential of PFAS incineration for releasing ozone-depleting chlorofluorocarbons and fluorinated greenhouse gases [34]. A study in Japan reported that after the thermal reactivation of granular activated carbon with adsorbed PFOA, PFOS, and PFHxS at 700 °C, a significant portion of the PFAS was converted to volatile species [86]. A US study reported measurable PFAS concentrations in the ash after incineration of sewage sludge [87]. Similarly, a 2021 study conducted a comparison study for the levels of PFAS in fly ash, bottom ash, and leachate from incineration plants. Higher levels of PFAS were observed in the leachate, when compared to the fly ash, and bottom ash. Although in lower concentrations, PFAS in the ashes were detected [82]. 1.12 Non-destructive treatment technologies 1.12.1 Adsorption Adsorption with granular activated carbon (GAC) has been one of the fastest and economically viable solutions to treat PFAS present in relatively pure water sources such as drinking water or groundwater [88]. The process is based on the physisorption and chemisorption of the target pollutants in the porous structure of the carbon. The mechanisms underlying the adsorption of Figure 1.9. Mechanism of competitive adsorption of long chain, short chain PFAS and organic matter (OM). Reprinted with permission from [15] 18 PFAS onto carbonaceous materials are electrostatic interactions, hydrogen bonding, hydrophobic interactions, and ligand exchange [89]. The factors influencing those interactions are ionic strength, ionic species, temperature, initial concentration of species, pH of the solution, coexisting matter, etc [89]. For the case of PFOA and PFOS, the hydrophobic effect and electrostatic interactions are the main driving forces for their adsorption [90]. Figure 1.9 shows the mechanism of competitive adsorption of long chain, short chain PFAS and organic matter. An increase in concentration of PFAS leads to the desorption of short chain PFAS, and prevents further adsorption of PFAS. In addition, the presence of organic matter leads to its dominance of the sorption sites on the sorbent’s surface and creates electrostatic interaction with the anion head of PFAS that repels them. Although organic matter attracts the hydrophobic tail of long chain PFAS, they do not diffuse within the sorbent [15]. The efficiency of activated carbon for the adsorption of long-chain PFAS such as PFOA and PFOS has been demonstrated in the past [91–94]. Typically, PFCAs are adsorbed faster than PFSAs due to the lower steric hindrance of the carboxylic group [15]. However, adsorption technologies are not selective for smaller chain PFAS [95]. This is because short chain PFAS are more hydrophilic when compared to the long-chain ones [15]. A study that evaluated the removal of 14 different PFAS using GAC in a continuous process found that the removal efficiency decreases with the increase of the carbon length of the perfluorinated compound [93] In addition, some short chain PFAS were desorbed after a period of time [93]. 1.12.2 Ion exchange Ion exchange is based on a dual mechanism: electrostatic attraction and adsorption. The structure of an IX resin possesses a backbone and exchange sites. The hydrophobic backbone is a neutral copolymer that adsorbs the hydrophobic part (fluorinated carbon chain) of the PFAS compound. The hydrophilic part (exchange sites) is functionalized with quaternary ammonium groups (positively charged) to attract the functional group of the PFAS molecules [96]. The selectivity of resins for PFAS increases with the length of the alkyl chains of the functional exchange group in the 19 exchange sites [96, 97]. However, the removal percentage of short chain PFAS decreases due to their low hydrophobicity. Additionally, the adsorption capacity decreases with the presence of co-contaminants such as chlorine and natural organic matter (NOM)[96]. 1.12.3 Foam fractionation Foam fractionation is a physical process that utilizes the surface-active properties of multiples PFAS to extract them from water. The process works by generating bubbles in the solution with the assistance of diffusers or venturi devices [66, 98]. Surface-active PFAS stick to the bubbles and travel to the air-liquid interface where they accumulate in the form of foam. The foam containing PFAS can be extracted by: 1) using a low vacuum pump to another reservoir where the bubbles burst due to the low pressure or, 2) creating an overflow to collect the foam containing PFAS. Some of the advantages of this technology are the low cost and small volumes recovered of solutions containing PFAS. However, some short-chain PFAS can not be removed. 1.13 Research motivation Among the technologies commercially available to remove PFAS from water, adsorption-based technologies have been considered the preferred alternative in terms of feasibility and scalability, given the emerging problem. However, the PFAS problematic has been only partially solved due to the fact that: i) the efficiency of adsorption-based technologies is limited to long-chain PFAS only; ii) adsorption-based technologies are not targeted for complex matrices such as landfill leachates or wastewater; iii) the spent adsorbent material is disposed in landfills or incinerated, creating an increasing PFAS accumulation cycle, and iv) the cost of carbon/resins is still relatively high for PFAS treatment and it increases with the complexity of the solution. There is the need to develop water treatment technologies that are able to break down the PFAS accumulation cycle and feasible to implement. During the last years, significant research efforts have been conducted on: 1) the development of destructive water treatment technologies able to degrade PFAS, and 2) the assessment of treatment trains that couple non-destructive with 20 destructive technologies to increase the feasibility of destructive technologies and decrease the cost of the overall treatment of PFAS. The current work addresses both of the previous needs. The destructive technology of choice is electrochemical oxidation, attributed to its high performance that has been demonstrated for the degradation of organic compounds. The electrochemical oxidation is targeted to complex matrices that include landfill leachates and IX still bottoms (waste from the IX process). The non-destructive technologies used are IX and a dieletrophoresis-enhanced adsorption. The former is used in a IX/EO treatment train for AFFF-impacted groundwater. The latter is used for the treatment of drinking water. 1.14 Research Objectives 1.14.1 General Objective The goal of this research was to develop and evaluate various treatment technologies to remove PFAS from multiple matrices. 1.14.2 Specific Objectives The specific objectives of this Ph.D. dissertation are as follows: 1. Identify the PFAAs present in landfill leachates from multiple sources in Grand Rapids, MI, and asses their electrochemical oxidation with a BDD flow-through cell. 2. Identify the target and suspect PFAS present in a representative landfill leachate and assess the electrochemical transformation of PFAAs and PFAA precursors over time. 3. Evaluate and optimize the electrochemical treatment of PFAAs from IX still bottoms at laboratory and semi-pilot scales. 4. 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Copyright 2021 Elsevier 31 2.1 Introduction Per-and polyfluoroalkyl substances (PFAS) are a group of synthetic chemicals widely used in multiple consumer products (e.g., tapestry, outdoor clothing, cleaning agents, non-stick cookware) and industrial processes (e.g., metal plating, fire-fighting foams, coatings, electronics) due to their unique surface-active properties and high chemical and thermal stability [1, 2]. An estimate of 3000 PFAS have been identified, from which perfluorooctanoic acid (PFOA) and perfluorooctanesulfonic acid (PFOS) are two of the most studied compounds [3]. PFAS have triggered attention due to their recalcitrant nature and bioaccumulative potential that leads to their accumulation in water, sediments, soils, wildlife, and the human body [3]. Their exposure and accumulation in the human body have been associated with multiple health effects (e.g., inmunotoxicity, neurotoxicity, testicular and kidney cancer) [4, 5]. As a result, the United States Environmental Protection Agency (USEPA) established a health advisory level (HAL) of 0.07 μg/L for the combined concentration of PFOA and PFOS in drinking water [6, 7]. Multiple PFAS end their life cycle in landfills as municipal solid waste, and their presence has been reported in landfill leachates in a wide range of concentrations [8, 9]. In 2013, for example, a range of 0.15−9.2 μg/L of PFOA was detected in 13 landfill leachate sites in the U.S [8]. A more recent study (2019), performed in Michigan U.S., estimated a daily flow of leachates from 32 landfills of over 1 million gallons with concentrations in the range of 16−3200 ng/L for PFOA and 9−960 ng/L for PFOS [10]. The concentration of PFAS in leachates is affected by various factors, including the heterogeneity of waste disposed, climate, waste age, and seasonal variability in infiltration [8, 9]. According to Lang et al., the most common PFAS present in landfill leachates in the U.S. are 5:3 fluorotelomer carboxylic acid (5:3 FTCA), perfluorohexanoic acid (PFHxA), perfluorobutanoic acid (PFBA), PFOA, 6:2 fluorotelomer carboxylic acid (6:2 FTCA), and perfluoropentanoic acid (PFPeA) [8]. Overall, PFAS ranging from C4-C8 chain length dominate the distribution profiles [11]. Wastewater treatment plants (WWTPs) receive landfill leachates as influents to be treated conventionally. Masoner et al. estimated that although landfill leachates accounted for only 32 1.7% of the total daily flow that goes into the studied WWTPs, the contribution of total PFAS corresponded to 18% of the total PFAS present in the influent of WWTPs [12]. In addition, previous studies have shown higher concentrations of PFAS in the effluent compared to the influent [8, 13]. This observation has been attributed to: 1) the non-biodegradability of PFAS; and 2) the fact that multiple polyfluoroalkyl substances (i.e., precursor compounds) can be further oxidized to perfluoroalkyl substances during biological treatment [8, 14]. Some of the precursors include fluorotelomer based substances (FTCAs), perfluoroalkyl sulfonamide derivatives (FASAAs), and polyfluoroalkyl phospahate esters (PAPs) [3, 9]. Additional treatment technologies, including adsorption with granular activated carbon (GAC) and membrane processes, i.e., nanofiltration (NF) and reverse osmosis (RO), have been proposed to treat PFAS in landfill leachates [15]. However, the complex composition of a landfill leachate makes GAC inefficient, while for the case of membrane processes, the concentrate containing PFAS require further treatment. Therefore, there is an urgent need for a destructive technology to degrade PFAS and break the accumulation cycle generated by other technologies. Electrochemical oxidation has shown to be a versatile destructive technology due to its capability to degrade a wide range of contaminants, operation at ambient temperature and pressure, and robust performance [16, 17]. Additionally, it does not require auxiliary chemicals and can be operated as a decentralized treatment option [17]. Multiple studies have been conducted to explore the electrochemical oxidation of PFAS in syn- thetic solutions and groundwater, showing promising results [18–20]. However, the effectiveness of the process in complex matrices, e.g., landfill leachate, membrane concentrates, and ion exchange regenerate solutions has been scarcely reported. Although the electrochemical oxidation of PFOA and PFOS in landfill leachates has recently been reported [16], multiple other PFAAs, commonly present in leachates, have only been identified but their oxidation has yet to be addressed. This study explores, for the first time, the electrochemical oxidation of multiple PFAAs in real landfill leachates using a boron-doped diamond (BDD) flow-through cell. The objectives of this work were to: i) evaluate and compare the degradation kinetics and energy consumption for 33 the electrochemical oxidation of two commonly studied PFAS (PFOA and PFOS) in a synthetic solution with a BDD flow-through cell; ii) assess the electrochemical oxidation of PFAAs in landfill leachates; and iii) determine the influence of leachates composition in the electrochemical oxidation of PFAAs in landfill leachates. 2.2 Materials and methods 2.2.1 Materials Perfluorooctanoic acid (PFOA, >97%) , heptadecafluorooctanesulfonic acid potassium salt (CF3 (CF2 )7 SO3 K, >98%), and perfluorobutanoic acid (PFBA, >98%), sodium sulfate (Na2 SO4 ) and sodium chloride (NaCl) were were purchased from Sigma Aldrich. 2.2.2 Landfill leachates Six leachate samples were collected from August 2019 to February 2020 from three different landfills in Michigan, USA. To maintain the confidentiality of sample locations, in this study, leachates were labeled as L1, L2, L3, L4, L5, and L6. The physico-chemical characterization of the samples is depicted in Tables 2.1 and 2.2. The leachates were collected in 20 L high density polyethylene (HDPE) containers, secured in coolers, and shipped to the Fraunhofer USA Center Midewest, Division for Coatings and Diamond Technologies at Michigan State University. Samples Table 2.1. Characterization of leachate samples Sample L1 L2 L3 L4 L5 L6 pH 7.94 8.26 7.95 7.79 7.89 8.07 Conductivity (mS/cm) 16.06 11.87 13.81 14.06 15.37 16.2 Chemical oxygen demand, COD (mg/L) 2205 1670 2560 2380 3000 5820 Total organic carbon, TOC (mg/L) 1320 910 940 1080 1100 1220 Nitrite (mg/L) 0 0 0 0 0.347 0.391 Nitrate (mg/L) 6.18 5.63 7.8 8.4 12.3 10.51 Ammonia, N-NH4+ (mg/L) 2210 1124 1676 2630 2200 2680 34 Table 2.2. Initial concentrations of PFAS, quantified in different leachate samples. All the values are shown in ng/L PFAS L1 L2 L3 L4 L5 L6 4:2 fluorotelomer sulfonate (4:2 FTS) BDL* BDL BDL BDL BDL BDL 6:2 fluorotelomer sulfonate (6:2 FTS) BDL BDL BDL BDL BDL BDL 8:2 fluorotelomer sulfonate (8:2 FTS) BDL BDL BDL BDL BDL BDL N-EtFOSAA BDL BDL BDL BDL BDL BDL N-MeFOSAA BDL BDL BDL BDL BDL BDL Perfluorobutanesulfonic acid (PFBS) 10000 4100 6600 5500 6100 2600 Perfluorobutanoic acid (PFBA) 3000 1800 2100 24000 1900 1800 Perfluorodecanesulfonic acid (PFDS) BDL BDL BDL BDL BDL BDL Perfluorodecanoic acid (PFDA) BDL BDL BDL BDL BDL BDL Perfluorododecanoic acid (PFDoA) BDL BDL BDL BDL BDL BDL Perfluoroheptanesulfonic Acid (PFHpS) BDL BDL BDL BDL BDL BDL Perfluoroheptanoic acid (PFHpA) 1400 1100 1100 1200 820 660 Perfluorohexanesulfonic acid (PFHxS) 1200 1800 1400 1400 800 510 Perfluorohexanoic acid (PFHxA) 5700 4000 4000 4900 3400 2800 Perfluorononanesulfonic acid (PFNS) BDL BDL BDL BDL BDL BDL Perfluorononanoic acid (PFNA) BDL BDL 2200 BDL BDL BDL Perfluorooctanesulfonamide (FOSA) BDL BDL BDL BDL BDL BDL Perfluorooctanesulfonic acid (PFOS) 2400 830 790 750 420 380 Perfluorooctanoic acid (PFOA) 3200 2200 4700 6700 1500 1200 Perfluoropentanesulfonic acid (PFPeS) BDL BDL BDL BDL 260 BDL Perfluoropentanoic acid (PFPeA) 2300 1500 1700 1900 1300 1100 Perfluorotetradecanoic acid (PFTeA) BDL BDL BDL BDL BDL BDL Perfluorotridecanoic acid (PFTriA) BDL BDL BDL BDL BDL BDL Perfluoroundecanoic acid (PFUnA) BDL BDL BDL BDL BDL BDL TOTAL 29200 17330 22390 46350 16500 11050 * BDL = Below detection limit. Detection limit for 4:2 FTS, 6:2 FTS, 8:2 FTS, NEtFOSAA, and NMeFOSAA corresponds to 2000 ng/L. Detection limit for PFBS, PFBA, PFDS, PFDA, PFDoA, PFHpS, PFHpA, PFHxS, PFHxA, PFNS, PFNA, FOSA, PFOS, PFOA, PFPeS, PFPeA, PFTeA, PFTriA, and PFUnA corresponds to 200 ng/L. were stored at 4 °C upon receipt. 35 Figure 2.1. Schematic of BDD flow-through cell 2.2.3 Electrochemical oxidation setup Experiments were performed at laboratory scale with a flow-through cell using niobium-supported BDD anodes and cathodes (Condias, Germany). The electrochemical cell was comprised of eight circular electrode packets. Each packet was formed by one BDD anode and two BDD cathodes, separated by an interelectrode distance of 2 mm. Each electrode was perforated with 60 holes (1/16" ID) to generate hydrodynamically turbulent conditions and allow the solution flow through the packets. The schematic of cell design is depicted in Figure 2.1. The total active anodic surface area was 33.6 cm2 . The cell was connected in parallel to two power supplies (BK precision 9130 B). A PVC tank was used as the reservoir/feed tank. Figure 2.2 shows a diagram of the experimental setup. A solution volume of V = 2 L was used to perform each experiment. The area to volume ratio (A/V) was 16.8 cm2 /L. Solutions were recirculated at a flow rate of 2 L/min using a peristaltic pump from the feed tank to the cell in a batch with recirculation set-up. All experiments were performed under galvanostatic conditions with the application of different current densities. The voltage ranged from 4.05 to 5.05 V for the lowest and the highest current density. In a typical experiment, 10 mL of leachate were collected from the reservoir tank every 2 h, transferred to polypropylene tubes, and stored in the refrigerator at 4◦ C until they were delivered for analysis. Additional parameters including chemical oxygen demand (COD), total organic carbon (TOC), and perchlorate (ClO4 – ) concentration were also monitored. No addition of electrolyte was required 36 Electrochemical cell Current Voltage Current Voltage Power supplies Peristaltic pump Reservoir tank Figure 2.2. Experimental set up for the electrochemical oxidation of PFAAs with a BDD flow- through cell. as the conductivity of the leachates was high enough to perform the experiments. The initial conductivity and pH had an average value of 14.6 ± 1.7 mS/cm and 7.9 ± 0.2, respectively. Control experiments to guarantee the absence of PFAS in all the components of the electrochemical reactor set-up were conducted. Pure water was recirculated through the system for one hour without the application of current. The final effluent was sent for PFAS analysis and showed no PFAS present. Two additional control experiments, without applied current, to test for PFAS losses (e.g., sorption or volatilization) not attributable to electrochemical treatment were performed with a synthetic solution containing PFOA and PFOS, and a leachate sample. Gas sampling was not considered in this work. 2.2.4 Analytical methods COD and TOC were determined using USEPA approved HACHTM standard methods. Anions were analyzed via ion chromatography (Dionex ICS-3000) using an ion exchange resin column IonPac AS20 (0.4 mm x 250 mm), based on Standard Methods 4110B. The pH and conductivity were measured with an SG23-B SevenGo DuoTM Series Portable Meter (Mettler Toledo). The temperature and flow rate were monitored using an in-house designed control system. PFAS analysis was performed following a modified EPA 537 method by Eurofins TestAmerica 37 (Sacramento, U.S.). Briefly, leachate samples were extracted using a solid phase extraction (SPE) cartridge with a Waters Oasis WAX 500 mg/6 cc column. PFAS were eluted from the cartridge with 0.3 % ammonium hydroxide/methanol. The final 80:20 methanol:water extracts were analyzed by LC/MS/MS with a Shimadzu CTO-20AC HPLC interfaced with a SCIEX 5500 Triple Quad MS. PFAS were separated from other components on a Phenomenex Gemini column (2.0 mm x 50 mm, 3 μm) with a solvent gradient program using 20 mM ammonium acetate/water and methanol. The details of the gradient method are provided in Table 2.3. The mass spectrometer detector was operated in electrospray (ESI) negative ion mode with a minimum of 10 scans/peak. The calibration standards used for PFAS detection are described in Table 4.4. For the quality assurance procedure, isotope dilution was used for the correction for analytical bias encountered in leachate samples. The isotope dilution analytes (IDA) consisted of carbon-13 labeled analogs, oxygen-18 labeled analogs, and deuterated analogs of the compounds of interest (details provided in Table 2.5) which were spiked into the samples at the time of extraction. Quantification by the internal standard method was employed for the IDA analytes/recoveries with the software Chrom Peak Review 2.1 using regression fit of r2 >0.90 and deviation <50%. The peak response was measured as the area of the peak. Seven calibration points in the range of 20 and 20000 ng/L were used for the quantification of the samples. Analyzed data were quantified if surrogate recovery was between 25 and 150%. The total identified PFAS precursors (TIP) (4:2 FTS, 6:2 FTS, 8:2 FTS, NEtFOSAA, NMeFOSAA) were below detection levels (<2000 ng/L) for all the samples in this work due to the dilution factor of the samples (100 ×). Table 2.3. Gradient solvent program for the HPLC Time (min) %A %B Flow rate (mL/min) 0 90 10 0.6 0.1 45 55 0.6 4.5 1 99 0.6 5.9 1 99 0.6 5.95 90 10 0.6 38 Table 2.4. Calibration standards used for PFAS detection Analyte LOD* MDL** Units % Recovery description limits 4:2 FTS 20.0 5.20 ng/L 79-139 6:2 FTS 20.0 2.00 ng/L 59-175 8:2 FTS 20.0 2.00 ng/L 75-135 N-EtFOSAA 20.0 1.90 ng/L 76-136 N-MeFOSAA 20.0 3.10 ng/L 76-136 PFBS 2.00 0.200 ng/L 67-127 PFBA 2.00 0.350 ng/L 76-136 PFDS 2.00 0.320 ng/L 71-131 PFDA 2.00 0.310 ng/L 76-136 PFDoA 2.00 0.550 ng/L 71-131 PFHpS 2.00 0.190 ng/L 76-136 PFHpA 2.00 0.250 ng/L 75-135 PFHxS 2.00 0.170 ng/L 59-119 PFHxA 2.00 0.580 ng/L 73-133 PFNS 2.00 0.160 ng/L 75-135 PFNA 2.00 0.270 ng/L 75-135 FOSA 2.00 0.350 ng/L 73-133 PFOS 2.00 0.540 ng/L 70-130 PFOA 2.00 0.850 ng/L 70-130 PFPeS 2.00 0.300 ng/L 66-126 PFPeA 2.00 0.490 ng/L 71-131 PFTeA 2.00 0.290 ng/L 70-130 PFTriA 2.00 1.30 ng/L 71-131 PFUnA 2.00 1.10 ng/L 68-128 * LOD = Limit of detection. ** MDL = Method detection limit. 2.3 Results and discussion 2.3.1 Performance of the BDD flow-through cell The electrochemical oxidation of two common PFAAs: PFOA and PFOS in a synthetic solution was evaluated with the flow-through cell. The solution consisted of 70 μg/L of PFOA and 70 39 Table 2.5. Surrogate used for PFAS detection Surrogate standards LOD* MDL** Units M2-4:2 FTS 50.0 25.0 ng/L M2-6:2 FTS 50.0 25.0 ng/L M2-8:2 FTS 50.0 25.0 ng/L d5-NEtFOSAA 50.0 25.0 ng/L d3-NMeFOSAA 50.0 25.0 ng/L 13 [ C3 ] PFBS 50.0 25.0 ng/L 13 [ C4 ] PFBA 50.0 25.0 ng/L [ 13C2 ] PFDA 50.0 25.0 ng/L 13 [ C2 ] PFDoA 50.0 25.0 ng/L [ 13C4 ] PFHpA 50.0 25.0 ng/L 13 [ C2 ] PFHxS 50.0 25.0 ng/L 13 [ C2 ] PFHxA 50.0 25.0 ng/L [ 13C5 ] PFNA 50.0 25.0 ng/L 13 [ C8 ] FOSA 50.0 25.0 ng/L [ 13C4 ] PFOA 50.0 25.0 ng/L 13 [ C2 ] PFTeDA 50.0 25.0 ng/L 13 [ C2 ] PFUnA 50.0 25.0 ng/L * LOD = Limit of detection ** MDL = Method detection limit. μg/L of PFOS dissolved in a 10 mM sodium sulfate (Na2 SO4 ) electrolyte. A current density of 50 mA/cm2 was applied during electrochemical treatment. Figure 2.3a shows the decrease in concentration of both PFOA and PFOS over time. Both species followed a pseudo-first order degradation kinetics (r2 = 0.9672 for PFOA and r2 = 0.9819 for PFOS) and the calculated values of the kinetic degradation constant for PFOA and PFOS were 2.19 × 10 – 2 and 3.99 × 10 – 2 min – 1, respectively. The degradation of (PFOA + PFOS) also followed a pseudo-first order Í degradation kinetics (r2 = 0.9873), with a rate constant of 2.63 × 10 –2 min –1 . Additionally, it has been shown that the degradation of long-chain PFAAs leads to the generation of shorter chain PFAAs (e.g., perfluoroheptanoic acid (PFHpA), PFHxA, perfluorohexanesulfonic acid (PFHxS), perfluoropentanoic acid (PFPeA), PFBA, and perfluorobutanesulfonic acid (PFBS)), as depicted in Figure 2.3b, that result from the cleavage of CF2 moieties [19, 21]. 40 (a) (b) 0 150 PFOA 1.0 −1 r2= 0.9672 PFBS PFHxA PFOS PFHxS PFHpA ln [Cf \ C0 ] −2 PFBA PFOS Total PFAAs (μg/L) PFPeA PFOA 0.8 −3 r2= 0.9819 100 −4 Cf \ C0 0.6 −5 0.5 1.0 1.5 2.0 Time (h) 0.4 50 0.2 0 0 0.5 1.0 1.5 2.0 0 0.5 1 1.5 2 Time (h) Time (h) Figure 2.3. (a) Decrease in concentration of PFOA and PFOS, and (b) byproducts of PFOA and PFOS oxidation over time during the electrochemical treatment of synthetic solutions with a BDD flow-through cell. Individual initial concentrations of PFOA and PFOS were 70 μg/L. Current density applied = 50 mA/cm2 . Solutions were prepared with 10 mM Na2 SO4 . An additional experiment with a BDD parallel-plate cell (shown in Figure 2.4) was performed for comparison using the same experimental conditions. The parallel-plate cell utilized a series of niobium-supported BDD rectangular parallel-plate electrodes (Condias, Germany). The cell had 2 anodes and 3 cathodes separated by 3 mm channels. The total active surface area of the anodes was Figure 2.4. Schematic of BDD parallel-plate cell 41 213.2 cm2 . The area to volume ratio (A/V) was 106.6 cm2 /L. A schematic of the cell is shown in Fig. 2.4. The electrochemical cell comparison experiments used the same experimental conditions for both the parallel-plate and flow-through cell. A normalized (with respect to current and treatment volume) pseudo-first order rate constant to describe the removal of PFOA and PFOS can be compared for different electrochemical systems [23]. Table 2.6 shows the normalized rate constants values for the electrochemical treatment of PFOA and PFOS attained in various studies. Results with the flow-through cell showed a higher normalized rate constant for the degradation of PFOA and PFOS when compared to the parallel- plate cell and other studies performed with similar conditions. The Electric Energy per Order (EE/O, Wh/L) required for the electrochemical oxidation of PFOA and PFOS was also considered for the evaluation of cell performance. As shown in Table 2.6, the EE/O values required for 1 log removal (90% degradation) of PFOA and PFOS with the BDD flow-through cell were 5 and 21 times lower, respectively, than the values obtained with the parallel-plate cell. The low EE/O is ascribed to multiple factors, including the geometry of the cell and the area to volume (A/V) ratio. The geometry of the cell heavily influences the diffusional limitations and energy losses of the process. The introduced flow-through cell increases the turbulence generated with the addition of multiple holes with different alignments on the surface area of the anodes and cathodes through which the solution flows, as shown in Figure 2.1, that Table 2.6. Comparison of current normalized rate constants and energy per order values for the electrochemical oxidation of PFOA and PFOS among various studies. All studies were performed in batch mode. All studies used a parallel-plate cell configuration, except for a in this work that used a flow-through cell Current Area/ Normalized rate constant EE/O density volume (min – 1 A – 1 L) (Wh/L) Anode Matrix (mA/cm2 ) (cm2 /L) PFOA PFOS PFOA PFOS Ref a –4 –4 BDD Synthetic solution flow-through 50 16.8 26.0 × 10 47.5 × 10 8.0 4.4 This work –4 –4 BDD Synthetic solution parallel-plate 50 106.6 10.5 × 10 5.0 × 10 42.6 89.9 This work –4 –4 BDD 10.6 mM Na2 SO4 + 0.05mM NaCl 50 152 2.6 × 10 0.8 × 10 180.0 500.0 [22] –4 –4 BDD Groundwater 25 40 3.8 × 10 1.4 × 10 160.1 438.1 [23] –4 –4 TSO 100 mM Na2 SO4 5 390 6.9 × 10 10.6 × 10 18.5 12.1 [24] –4 –4 TiRuO2 Groundwater + 500 mg/L Na2 SO4 20 50 7.3 × 10 6.5 × 10 68.3 76.7 [25] 42 leads to an enhancement in the mass transfer coefficient km . The value of km was electrochemically –5 –4 determined as described elsewhere [26] and corresponded to 1.22 × 10 and 1.25 × 10 m s–1 for the parallel-plate cell and the flow-through cell, respectively. Moreover, energy losses can be reduced with the minimization of the interelectrode distance that is responsible for the ohmic drop of the cell [27], which was the case for the flow-through cell. The A/V ratio corresponds to the area of electrodes used with respect to the treated water volume. An optimization of this parameter allows for reduction of capital costs of the technology which is highly dependent on the electrode area used. In this regard, as shown in Table 1, the present work with the flow-through cell for the degradation of PFOA and PFOS used the lowest A/V ratio reported to date and nonetheless provided high degradation rate constants and low EE/O values. After evaluating the performance with synthetic solutions, the BDD flow-through cell was tested with real landfill leachates. The results are presented in the following subsections. 2.3.2 Influence of current density on the electrochemical treatment of PFAAs in landfill leachates The influence of the current density on the degradation of multiple PFAAs present in leachate L1 was evaluated. This leachate was spiked with 25 μg/L of PFOA and 15 μg/L of PFOS to increase their concentration as these two compounds are the ones currently regulated. The detected concentration of PFAAs in the spiked leachates from L1 ranged from high to low in the following order: PFOA, PFOS, PFBS, PFHxA, PFBA, PFPeA, PFHpA, and PFHxS. Preliminary experiments (data not shown) were conducted to determine the current density range in which electrochemical oxidation of PFAAs occurs for the A/V used (16.8 cm2 /L). Current densities lower than 50 mA/cm2 led to an increase in PFOA and PFOS, and only current densities equal to or greater than 50 mA/cm2 allowed for their decrease in concentration. Therefore, a range from 50 to 200 mA/cm2 was selected for the following experiments. Figure 2.5 shows the concentration of detected PFAAs over time during the electrochemical treatment of L1 with 50, 100, 150, and 200 mA/cm2 . In general, the increase in current density 43 (a) (b) 120 120 PFPeA PFHxA PFOS PFHpA 100 PFOA PFHxS PFBS PFBA 100 Total PFAAs (μg/L) Total PFAAs (μg/L) 80 80 60 60 40 40 20 20 0 0 0 2 4 6 8 0 2 4 6 8 Time (h) Time (h) (c) (d) 120 120 100 100 Total PFAAs (μg/L) Total PFAAs (μg/L) 80 80 60 60 40 40 20 20 0 0 0 2 4 6 8 0 2 4 6 8 Time (h) Time (h) Figure 2.5. Concentration of total PFAAs over time during the electrochemical oxidation of leachate L1 with a BDD flow-through cell. The applied current densities were: (a) 50 mA/cm2 , (b) 100 mA/cm2 , (c) 150 mA/cm2 , and (d) 200 mA/cm2 . Samples were spiked with [PFOA]0 ≈ 25 μg/L and [PFOS]0 ≈ 15 μg/L. allowed for a faster removal of total PFAAs. In the cases of 50 and 100 mA/cm2 , the concentration of total PFAAs after 8 h of electrochemical treatment was higher than its initial concentration. Nevertheless, most of the final concentration corresponded to PFBA and represented 59.0 and 67.4% of the final concentration of total PFAAs in L1 treated with 50 and 100 mA/cm2 , respectively. Conversely, current densities of 150 and 200 mA/cm2 led to a decrease in total PFAAs after 8 h of treatment and the final concentration of PFBA was minimized with 200 mA/cm2 . The removal of perfluoroalkyl carboxylic acids (PFCAs) had a strong dependence on the current density applied, 44 contrary to perfluoroalkyl sulfonic acids (PFSAs), which removal was independent, as shown in Figure 2.6. This observation implies that PFSAs: 1) are easier to degrade than PFCAs and 2) have a degradation mechanism that supports conversion to PFCAs. This finding is supported by previous research, which suggested that the degradation of PFSAs initiates with the cleavage of the SO3 – group from the terminal carbon and formation of perfluoroalkyl radicals, followed by multiple chain reactions that lead to the formation of short-chain PFCAs [16, 28]. This transformation mechanism was also observed for the electrochemical oxidation of precursor compounds such as 6:2 FTSA [29]. Additionally, it has been suggested that the reaction of precursor compounds with •OH leads to the generation of a mixture of PFCAs of varying carbon chain length [30]. The removal of individual Í chains was also compared. For the case of (PFOA + PFOS), after 2 h of treatment, a removal percentage higher than 90% was reached with 100, 150, and 200 mA/cm2 , whereas 50 mA/cm2 allowed for only 72% removal. A treatment time of 8 h was required to reach a removal percentage of 90% with 50 mA/cm2 . The latter followed pseudo-first order degradation kinetics (𝑘 = 4.37 × –3 10 min – 1 and r2 = 0.7998), six times slower than the degradation with the same current density in synthetic solutions. The slower degradation was attributed to the presence of multiple other co- contaminants in the leachate that compete for oxidation with PFAAs. In general, the concentration of PFAAs in a leachate is low relative to other components present in the matrix (e.g., dissolved (a) (b) 2.4 1.0 50 mA/cm2 50 mA/cm2 100 mA/cm2 100 mA/cm2 150 mA/cm2 2.0 150 mA/cm2 200 mA/cm2 200 mA/cm2 CPFSAs \ CPFSAso CPFCAs \ CPFCAso 0.8 1.6 0.6 1.2 0.4 0.8 0 2 4 6 8 0 2 4 6 8 Time (h) Time (h) Figure 2.6. Concentration of (a) PFSAs and (b) PFCAs during the electrochemical oxidation of leachate L1 with multiple current densities. 45 organic matter (DOM), multiple other xenobiotic organic compounds (XOCs), heavy metals, and inorganic salts) [3]. Therefore, multiple other co-contaminants are oxidized simultaneously. The removal of shorter-chain PFAAs (4 ≤ C ≤ 8) was also dependent on the current density. For the shorter chain PFCAs, the concentration of PFHpA and PFHxA increased over time (8 h) with 50 mA/cm2 by a factor of 2.0 and 1.7, respectively, and decreased with 150 and 200 mA/cm2 . The concentration of PFPeA and PFBA increased with all current densities, yet, the final concentration after treatment with 200 mA/cm2 was lower than with 50 mA/cm2 . For the shorter chain PFSAs, the concentration of PFHxS decreased with all the current densities, leading to non-detect values after 2 h with 150 and 200 mA/cm2 . Conversely, the concentration of PFBS increased over time with all the current densities applied. However, the generation of PFBS was at least 14 times less than PFBA. The increment of PFPeA, PFBA, and PFBS is in agreement with the observations in previous studies. For instance, Gomez-Ruiz et al. showed that the increase in current density for the degradation of 6:2 FTSA in industrial wastewater decreased the total concentration of PFCAs, but increasing trends for PFHxA and PFBA were observed [31]. Trautmann et al. reported the increase of PFBA, PFPeA, and PFHxA after 18 h of electrooxidation of simulated groundwater containing various PFAS [32]. The common pattern of decreasing concentrations of longer chains and increasing concentra- tions of shorter chains follows the previously proposed PFAS unzipping mechanism [19, 21]. In this mechanism, PFAS are activated by a direct electron transfer to form perfluoroalkyl radicals, which then react with •OH in a series of chain reactions to form shorter-chain PFAS. Although it has been shown that PFAS are inert to radical attack, the perfluoroalkyl radicals are vulnerable to • OH [33, 34]. Both direct and indirect oxidation occur concurrently until complete mineralization is achieved [33]. For the present work, the complexity of a leachate matrix likely slows down the mineralization of short-chain PFAAs. The increase in concentration of PFBA in all the cases can be attributed to two coexisting processes: degradation of longer chains [23], and transformation of potential precursor compounds 46 (a) (b) 2.5 2.5 Fraction of Molar F in PFCAs Fraction of Molar F in PFCAs PFPeA PFPeA PFOA PFOA PFHxA PFHxA PFHpA PFHpA 2.0 PFBA 2.0 PFBA 1.5 1.5 1.0 1.0 0.5 0.5 0 0 0 2 4 6 8 0 2 4 6 8 Time (h) Time (h) (c) (d) 2.5 2.5 Fraction of Molar F in PFCAs Fraction of Molar F in PFCAs PFPeA PFPeA PFOA PFOA PFHxA PFHxA PFHpA PFHpA 2.0 PFBA 2.0 PFBA 1.5 1.5 1.0 1.0 0.5 0.5 0 0 0 2 4 6 8 0 2 4 6 8 Time (h) Time (h) Figure 2.7. Fraction of molar F relative to 𝑡 = 0 in PFCAs during the electrochemical oxidation of leachate L1 with (a) 50 mA/cm2 , (b) 100 mA/cm2 , (c) 150 mA/cm2 , and (d) 200 mA/cm2 . present in leachates [13, 35] that ultimately were oxidized to PFBA. The latter was confirmed with a mass balance of the organic fluorine in PFAAs as shown in Figures 2.7 and 2.8. The fraction of molar F relative to 𝑡 = 0 in PFCAs and PFSAs during the electrochemical oxidation of PFAAs in leachates with different current densities is depicted. For the experiments performed with 50, 100, and 150 mA/cm2 , the organic fluorine for PFCAs increased by a factor of 2.0, 1.8 and 1.3 after 8 h of treatment, respectively. This excess of fluorine was likely generated from the oxidation of precursors that were oxidized to PFCAs. This assumption is based on the identification of precursor compounds in previous studies. For instance, Lang et al. reported the presence of precursor compounds in concentrations higher than the limit of quantification (LOQ) for more than 47 (a) (b) 1.2 1.2 Fraction of Molar F in PFSAs Fraction of Molar F in PFSAs PFOS PFOS PFHxS PFHxS 1.0 PFHpS PFBS 1.0 PFHpS PFBS 0.8 0.8 0.6 0.6 0.4 0.4 0.2 0.2 0 0 0 2 4 6 8 0 2 4 6 8 Time (h) Time (h) (c) (d) 1.2 1.2 Fraction of Molar F in PFSAs Fraction of Molar F in PFSAs PFOS PFOS PFHxS PFHxS 1.0 PFHpS PFBS 1.0 PFHpS PFBS 0.8 0.8 0.6 0.6 0.4 0.4 0.2 0.2 0 0 0 2 4 6 8 0 2 4 6 8 Time (h) Time (h) Figure 2.8. Fraction of molar F relative to 𝑡 = 0 in PFSAs during the electrochemical oxidation of leachate L1 with (a) 50 mA/cm2 , (b) 100 mA/cm2 , (c) 150 mA/cm2 , and (d) 200 mA/cm2 . 50% of the leachates (95 samples) analyzed from 18 landfills in the U.S [8]. In addition, it has been shown that the transformation of some precursors can lead to the generation of PFOA and PFOS [36, 37]. The identification and study of the transformation of precursors compounds will be presented in future work. Likewise, it has been shown that more hydrophobic PFAAs are easier to degrade. Among all the PFAAs studied in this work, PFBA was the least hydrophobic, hence, it had the slowest degradation kinetics, as shown in a previous study [24]. With this precedent, a generation rate higher than the degradation rate presumably led to the increase in PFBA concentration during the electrochemical oxidation process. An additional experiment was performed using synthetic 48 1.0 Na2SO4 NaCl 0.8 C / C0 0.6 0.4 0.2 0 0 2 4 6 Time (h) Figure 2.9. Electrochemical degradation of PFBA (1 mg/L) with a current density of 150 mA/cm2 using Na2 SO4 and NaCl as Chapter2porting electrolytes. 1.0 0.8 COD/COD0 0.6 0.4 50 mA/cm2 0.2 100 mA/cm2 150 mA/cm2 200 mA/cm2 0 0 2 4 6 8 Time (h) Figure 2.10. Evolution of concentration of COD with respect to t=0 over time during the electro- chemical oxidation of leachate L1 with multiple current densities. [PFOA]0 ≈ 28 μg L – 1 ; [PFOS]0 ≈ 18 μg L – 1 . solutions to verify the capability of BDD to oxidize PFBA and is shown in Figure 2.9. PFBA was degraded by 98.4 and 80.8% with 150 mA/cm2 after 6h of treatment with Na2 SO4 and NaCl as supporting electrolytes, respectively. The low levels of inorganic fluoride (LOQ of 1 mg/L), low concentrations of total PFAAs (low μg/L range), and the complexity of the matrix did not allow us to quantify the fluoride generation. Additional parameters that influence the electrochemical oxidation process were evaluated in this set of experiments to evaluate the overall oxidation process of leachates. These included chemical oxygen demand (COD) and total organic carbon (TOC). Figure 2.10 shows the COD evolution over time during the electrochemical oxidation of landfill 49 leachates. The decrease in COD followed a first order kinetics for all the current densities with degradation constant rates (𝑘, min – 1 ) proportional to the applied current density (Table 2.7), as observed in a previous study [38]. A maximum COD removal of 86% was reached after 8 h of treatment with 200 mA/cm2 . The TOC removal was also quantified and corresponded to 17, 42, 68, and 73% after 8 h of treatment with 50, 100, 150, and 200 mA/cm2 , respectively. These values suggest incomplete mineralization for all the applied conditions, which is in agreement with the generation of multiple observed short chain PFAAs, and additional non-target compounds that remained in the final treated solution. 2.3.3 Electrochemical treatment of various leachates In the next part of this study, the degradation of PFAAs was evaluated for 5 different leachates: L2, L3, L4, L5, and L6. The individual characterization of each leachate is depicted in Tables 2.1 and 2.2. None of the leachates were spiked. The goal of this set of experiments was to evaluate the influence of the leachate characteristics on the electrochemical degradation of PFAAs and to determine the existing correlations between variables. The experiments were performed with a current density of 150 mA/cm2 applied for 6 h. The latter was chosen over 200 mA/cm2 as it showed to be sufficient to oxidize total PFAAs and required lower energy consumption. Figure 2.11a depicts the initial concentrations of individual PFAAs corresponding to 5 different leachate samples. PFBA, PFBS, PFHpA, PFHxA, PFHxS, PFOA, PFOS, and PFPeA were found Table 2.7. Values of kinetic rate constants for COD evolution during the electrochemical oxidation of leachate L1 with multiple current densities Current density k 2 –2 –1 r2 (mA/cm ) (10 min ) 50 1.43 0.9922 100 2.77 0.9917 150 3.15 0.9926 200 4.35 0.9909 50 (a) (b) 104 100 Concentration (ng/L) Removal efficiency (%) 50 0 3 10 −50 −100 PFOA PFOS PFHpA PFHxA PFHxS PFPeA PFBA PFBS PFOA PFOS PFHpA PFHxA PFHxS PFPeA PFBA PFBS PFAAs PFAAs Figure 2.11. Box and whisker plot for: (a) initial concentrations of PFAAs detected in five different leachate samples (L2−L6), and (b) removal efficiency (%) of leachates L2−L6 after 2 h of electrochemical oxidation with an applied current density of 150 mA/cm2 . Removal efficiency is between +100 and −100% . Ends of the boxes represent the first and third quartiles, horizontal line inside the box represent the median, whiskers represent minimum and maximum values. Samples were not spiked. in all the samples. The total PFAAs concentration varied between 11.1 to 24.8 μg/L (mean 18.4 ± Í Í 4.8 μg/L). The PFCAs ranged from 7.6 to 17.1 μg/L (mean 11.6 ± 3.4 μg/L) and the PFSAs ranged from 3.5 to 8.8 μg/L (mean 6.8 ± 1.8 μg/L), indicating that PFCAs were the dominant species, which has also been observed in previous studies [3, 9]. Initial concentrations of PFCAs and PFSAs for each landfill leachate are shown in Table 2.8. Initial concentrations of individual PFAAs ranged from 0.6 to 6.7 μg/L (Table 2.2). The mean concentration was the highest for PFBS Table 2.8. Initial concentration of PFCAs, PFSAs, and total PFAS of the leachates treated in this study PFSAs PFCAs TOTAL Landfill (ug/L) (ug/L) PFAS (ug/L) L2 6.73 10.60 17.33 L3 8.79 13.60 22.39 L4 7.65 17.10 24.75 L5 7.58 8.92 16.50 L6 3.49 7.56 11.05 51 (4.9 ± 1.5 μg/L) and the lowest for PFOS (0.6 ± 0.2μg/L). Low concentrations of PFOS in landfill leachates relative to other PFAAs have been also reported elsewhere [39, 40]. For other PFAAs, the mean concentrations were from high to low: PFHxA (3.8 ±0.7 μg/L), PFOA (3.3 ± 2.1 μg/L), PFBA (2.0 ± 0.2 μg/L), PFPeA (1.5 ± 0.3 μg/L), PFHxS (1.2 ± 0.5 μg/L), Í and PFHpA (1.0 ± 0.2 μg/L). The concentration of (PFOA + PFOS) (mean 3.9 ± 2.5 μg/L) was between 23 and 106 times above the USEPA HAL of 70 ng/L in the samples tested. Other unregulated PFAAs: PFHxS, PFHpA, and PFBS, were detected at mean concentrations of 39, 98, and 55 times higher than their minimum reporting levels of 0.03, 0.01, and 0.09 μg/L, respectively, based on the USEPA’s third Unregulated Contaminant Monitoring Rule (UCMR3). Figure 2.11b shows the removal percentage of individual PFAAs after 2 h of treatment where PFOS reached non-detect levels (mean 100%) and PFOA reached a degradation percentage higher than 97% (mean 97 ± 4%) for all samples. For other PFAAs, the average removal efficiencies were 88 ± 17, 29 ± 13, and 96 ± 7% for PFHpA, PFHxA, and PFHxS, respectively. Negative removal (increasing concentration) was observed for PFPeA, PFBA, and PFBS, with values of −16 ± 31, −181 ± 235, and −3 ± 22%, respectively. Increasing the treatment time up to 6h enhanced the removal efficiency of PFHxA and PFBS, but led to a higher negative removal of PFPeA and PFBA. As stated before, the increasing concentration of short-chain PFAAs is a result of the degradation of longer chains, preferential conversion of PFSAs to PFCAs, and possible transformation of precursor compounds [23, 31]. The PFAAs compound with the highest concentration after 6 h of treatment was PFBA for leachates L2, L3, L4, and PFHxA for leachates L5 and L6. The concentration of PFAAs over time during the electrochemical treatment of each leachate treated is depicted in Figure 2.12. The combination of positive and negative removal for individual PFAAs led to a negative total PFAAs removal for L2 (−138,6%) and L3 (−64.7%), and a positive total PFAAs removal for L4 (48.3%), L5 (67.6%), and L6 (73.5%). A Pearson’s correlation analysis was performed to determine the correlation between the leachates characteristics and total PFAAs removal. A positive significant correlation was found be- tween initial TOC and total PFAAs removal (r = 0.92, p = 0.028). COD was moderately correlated 52 (a) (b) 50 50 PFPeA PFPeA PFOS PFOS PFOA PFOA 40 PFHxS 40 PFHxS Total PFAAs (μg/L) Total PFAAs (μg/L) PFHxA PFHxA PFHpA PFHpA PFBS PFBS PFBA PFBA 30 30 20 20 10 10 0 0 0 2 4 6 0 2 4 6 Time (h) Time (h) (c) (d) 50 50 PFPeA PFPeA PFOS PFOS PFOA PFOA 40 PFHxS 40 PFHxS Total PFAAs (μg/L) Total PFAAs (μg/L) PFHxA PFHxA PFHpA PFHpA PFBS PFBS PFBA PFBA 30 30 20 20 10 10 0 0 0 2 4 6 0 2 4 6 Time (h) Time (h) (e) 50 PFPeA PFOS PFOA 40 PFHxS Total PFAAs (μg/L) PFHxA PFHpA PFBS PFBA 30 20 10 0 0 2 4 6 Time (h) Figure 2.12. Concentration of total PFAAs over time during the electrochemical oxidation of multiple landfill leachates with a BDD flow-through cell. The landfills correspond to: (a) L2, (b) L3, (c) L4, (d) L5, and (e) L6. The applied current density was 150 mA/cm2 . Samples were not spiked. 53 (r = 0.65 , p = 0.238; although statistically insignificant). The initial concentration of total PFAAs was negatively and poorly correlated (statistically insignificant) with with the total PFAAs removal (r = −0.23, p = 0.706). The significant correlation of initial TOC and total PFAAs removal shows a dependency between variables where the percentage of PFAAs removal is affected by the level of carbon containing compounds. With this in consideration, electrochemical oxidation of PFAAs in landfill leachates should be applied as a decentralized treatment option at the point of source, as the variability of composition of different leachates determines the necessary treatment time to achieve higher removal efficiencies. Finally, the energy consumption for the electrochemical oxidation of PFAAs in landfill leachates was 28 and 82 Wh/L for 2 h and 6h of treatment, respectively. Once again, only 2 h (28 Wh/L) were necessary to reach a removal percentage higher than 90% of both PFOA and PFOS in all leachates treated. Although the process was applied to such a complex matrix, the energy consumption was still lower than the values reported for leachates in previous research [16]. 2.3.4 Perchlorate generation in leachates ClO4 – is a well-known byproduct of electrochemical oxidation that results from the oxidation of chlorinated compounds. The non-selective nature of electrochemical technologies leads to the oxidation of not only the target pollutants, but non-targeted compounds, which can be either beneficial or detrimental depending on the matrix and desired compounds to be removed. Chloride (Cl – ) is one of the components with the highest concentration in wastewater and landfill leachates [31]. In this work, the studied leachates presented an average initial Cl – concentration of 3026 ± 421 mg/L. It has been shown that the presence of high concentrations of Cl – in landfill leachates leads to the generation of reactive chlorine (Cl2 ), followed by its hydrolytic disproportionation to form hydrochlorous acid (HOCl) and hypochlorite ions (OCl – ). The foregoing contribute to the indirect oxidation of organic pollutants [17, 38] and it has been shown that reactive chlorine oxidizes the non-fluorinated head groups of PFAAs precursors via 54 50 mA/cm2 12 100 mA/cm2 150 mA/cm2 Perchlorate (mM) 200 mA/cm2 8 4 0 0 2 4 6 8 Time (h) Figure 2.13. Concentration of perchlorate over time during the electrochemical oxidation of landfill leachates with multiple current densities. [PFOA]0 ≈ 28 μg/L; [PFOS]0 ≈ 18 μg/L. Samples correspond to leachate L1. indirect oxidation mechanisms [34]. However, Cl – also can act as scavenger of •OH radicals to form products with higher oxidation states, including chlorate (ClO3 – ) and ClO4 – , [17, 33] the latter being the most common byproduct of electrochemical oxidation [17]. Figure 2.13 shows the concentration of ClO4 – over time during the electrochemical treatment of PFAAs with different current densities. The kinetics for ClO4 – followed a zero-order generation rate (shown in Table 2.9), which was independent on the initial concentration of Cl – , but dependent on the applied current density. Although avoiding the presence of Cl – in leachates may be difficult, the generation of ClO4 – can be diminished by using low current densities (as shown in Fig 2.13), Table 2.9. Values of zero order kinetic rate constants for perchlorate generation during the electrochemical oxidation of leachate L1 with multiple current densities Current density k 2 r2 (mA/cm ) mg/(L · min) 50 0.536 0.9445 100 0.759 0.9930 150 2.049 0.9805 200 2.541 0.9894 55 shorter treatment times, [41] or quenching the production of HOCl, and OCl – [33]. Additionally, biological treatment has been proposed as one of the alternatives to treat ClO4 – after electrochemical oxidation [22, 42]. For instance, Schaefer et al. yielded a 3−log decrease in ClO4 – levels generated during electrochemical oxidation using biological reduction [22]. All of these alternatives will have to be evaluated in future research to determine their implications. 2.4 Conclusions The results presented herein introduced a higher performance cell (flow-through) for the electro- chemical oxidation of PFAAs, allowing for lower energy consumption and enhanced mass transfer than a conventional parallel-plate cell. The concentrations of PFAAs in six different leachates from three landfill leachates in Michigan ranged from 10 2 to 10 4ng/L. PFCAs were in higher concentrations than PFSAs. The compounds PFOA and PFBS were identified as the PFAAs with the highest concentrations. Subsequently, a boron-doped diamond (BDD) flow-through cell was used to evaluate the electrochemical oxidation of PFAAs. The performance of the flow-through cell was assessed and compared with synthetic solutions for the oxidation of PFOA and PFOS. The results showed 6-times slower degradation rate for the electrochemical oxidation of PFOA and PFOS in landfill leachates when compared to synthetic solutions. The electrochemical oxidation of various leachates with a current density of 150 mA/cm2 led to a total PFAAs removal that ranged from -138.6 to 73.5%. Non-detect levels and degradation percentages higher than 97% for the oxidation of PFOS and PFOA respectively, were reached for all the leachates electrochemically treated. Although high removal efficiencies for long chain PFAAs, including PFOA and PFOS, were achieved for all samples, the degradation percentage of short-chain PFAAs, in particular PFBA, PFBS, and PFPeA, was lower and remains a challenge. A further study of the precursors influence and transformation needs to be considered in order to gain a better understanding of their implications for the electrochemical treatment of landfill leachates. Pretreatment technologies, aiming to preconcentrate PFAAs in leachates, may improve the PFAAs degradation efficiency by reducing the treatment volume and eliminating some of the 56 competitive species from the matrix. In addition, optimizing cell geometries could further enhance PFAAs degradation rates. With the previous appropriately combined, electrochemical oxidation could contribute to multiple integrated treatment processes, aiming to destroy PFAS. 57 BIBLIOGRAPHY 58 BIBLIOGRAPHY [1] H. Yan, I. T. Cousins, C. Zhang, Q. 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Hydrol. 89 (2007). 62 CHAPTER 3 ELECTROCHEMICAL TRANSFORMATIONS OF PERFLUOROALKYL ACID (PFAA) PRECURSORS AND PFAAS IN LANDFILL LEACHATES This chapter was reprinted with permission from Maldonado, V. Y.; Schwichtenberg, G.; Schmokel, S.; Witt, S.; and Field, J. Electrochemical Transformations of Perfluoroalkyl Acid (PFAA) Precur- sors and PFAAs in Landfill Leachates. J. Environ. Sci. Technol. Water., 2022, 2, 4, 624–634 https://pubs.acs.org/doi/10.1021/acsestwater.1c00479 Copyright 2022 American Chemical Society 63 3.1 Introduction Per-and polyfluoroalkyl substances (PFAS) are a group of fluorine-based compounds produced since the 1940s [1]. Their hydrophobic and oleophobic nature, in addition to their exceptional chemical and thermal stability [2–4] support a wide spectrum of industrial and consumer applications (e.g., firefighting foams, household products, food coatings, textiles) [4, 5]. The discharge of PFAS- containing products to the environment leads to i) the proliferation of PFAS in multiple water sources (e.g., surface water, groundwater, wastewater); ii) the transformation of perfluoroalkyl acid (PFAA) precursor compounds into two classes of recalcitrant PFAS, perfluoroalkyl carboxylic acids (PFCAs) and perfluoroalkyl sulfonic acids (PFSAs); and iii) the bioaccumulation of PFCAs and PFSAs in wildlife [6] and human-sera [7, 8]. Exposure to PFAS and accumulation in the human body are linked to multiple adverse health effects [9, 10]. Thus, PFAS are classified as emerging toxic compounds, and guidelines to control their release to the environment have been established worldwide [3]. Landfills are considered the final disposal point for products and wastes from residential, commercial, and industrial sources [11]. The composition of landfill leachates is complex and includes products of anaerobic decomposition, high concentrations of ammonia, chemical oxygen demand, salts, trace levels of metals, and xenobiotic organic compounds such as pharmaceuticals, pesticides, and of immediate interest, PFAS [11, 12]. Typically, leachates are collected and sent to wastewater treatment plants (WWTPs) for treatment [13]. Masoner et al. estimated that the contribution of total PFAS load from landfill leachates corresponded to 18% of the total PFAS load in the influent of WWTPs [14], rendering landfill leachates a significant secondary source of PFAS to the environment [15]. Concentrations of PFAS ranging from ng/L to μg/L are detected in landfill leachates worldwide, with PFCAs and fluorotelomer carboxylates (FTCAs) accounting for the classes with the highest concentration [15–17]. The broad concentration range is attributed to the heterogeneity of waste, landfill age, and climate conditions [14, 15]. Some of the most common PFAS classes detected in leachates include perfluroalkyl acids (PFAAs) (e.g., PFCAs, PFSAs), and multiple PFAA precursors such as: saturated (n:2 FTCA, n:3 FTCA) and unsaturated (n:2 UFTCA) 64 fluorotelomer carboxylic acids, fluorotelomer sulfonates (n:2 FTSs), perfluoroalkyl sulfonamide- based substances (perfluoroalkane sulfonamides (FASAs), perfluoroalkane sulfonamido acetic acids (FASAAs), and N-alkyl FASAAs), and polyfluoroalkyl phosphate esters (PAPs) [18, 19]. The PFAA precursors are ultimately transformed to PFAAs [12, 20–22]. Some of the effective technologies to remove PFAS from drinking water include granular activated carbon (GAC), ion-exchange resins (IX), nanofiltration (NF), and reverse osmosis (RO) [23]. However, GAC and IX are ineffective in treating PFAS from landfill leachates due to the high complexity of the matrix [24]. Furthermore, although NF and RO are able to concentrate PFAS from leachates in small volumes, further treatment to destroy PFAS is still required [11]. Electrochemical oxidation has shown its destructive potential for PFAS in multiple matrices, including synthetic solutions [25, 26], groundwater [27, 28], wastewater [29], and landfill leachates [30, 31]. Our previous work demonstrated the electrochemical oxidation of only eight PFAAs (C4−C8 PFCAs and C4, C6 and C8 PFSAs) in various landfill leachates, and provided evidence that the substantial increase of PFAAs, in particular perfluorobutanoic acid (PFBA), during the electrochemical treatment, was attributed to the transformation of non-identified PFAA precursors and the degradation of longer chain PFAAs [31]. Albeit the electrochemical transformation of PFAA precursors has been studied in groundwater [32], the intermediate and final products of precursors have not been reported for landfill leachates. The goals of this study were to: i) identify the target and suspect (Level 2b and 4) PFAS present in a landfill leachate and ii) assess the electrochemical transformation of target and suspect PFAS over time. In particular, intermediate and final products observed during the electrochemical treatment were fitted into previously reported PFAA transformation pathways of the most representative classes. Findings of this work include multiple PFAS reported for the first time in landfill leachates; and evidence for previously reported PFAA precursor transformation pathways in the studied leachate during the electrochemical treatment. The implications of the electrochemical treatment of leachates with high concentrations of PFAA precursors are discussed. 65 Table 3.1. Surrogate standards used for target PFAS analysis Chemical name Abbreviation Perfluoro-n-[ 13C4 ] butanoic acid MPFBA 13 Perfluoro-n-[3,4,5- C3 ] pentanoic acid M3PFPeA 13 Perfluoro-n-[1,2- C2 ] hexanoic acid MPFHxA Pefluoro-n-[1,2,3,4- 13C4 ] heptanoic acid M4PFHpA 13 Perfluoro-n-[1,2- C2 ] octanoic acid M2PFOA Perfluoro-n-[1,2,3,4- 13C4 ] octanoic acid MPFOA 13 Perfluoro-n-[1,2,3,4,5- C5 ] nonanoic acid MPFNA 13 Perfluoro-n-[1,2- C2 ] decanoic acid MPFDA Perfluoro-n-[1,2- 13C2 ] undecanoic acid MPFUdA 13 Perfluoro-n-[1,2- C2 ] dodecanoic acid MPFDoA Perfluoro-n-[1,2- 13C2 ] tridecanoic acid M2PFTeDA 13 Perfluoro-n-[1,2- C2 ] hexadecanoic acid M2PFHxDA 13 Sodium perfluoro-1-[2,3,4- C3 ]-butanesulfonate M3PFBS Sodium perfluoro-1-hexane[ 18O2 ] sulfonate MPFHxS 13 Sodium perfluoro-1-[1,2,3,4- C4 ]-octanesulfonate MPFOS Sodium perfluoro-1-[ 13C8 ]-octanesulfonate M8PFOS 18 Sodium perfluoro-1-hexane[ O2 ] sulfonate MPFHxS 13 Perfluoro-1-[ C8 ]octanesulfonamide M8FOSA-I N-methylprefluoro-1-octane sulfonamide d7-N-MeFOSA-M N-ethyl-d5 -perfluoro-1-octanesulfonamide d-N-EtFOSA-M N-methyl-d3 -perfluoro-1-octane-sulfonamidoacetic acid d3-N-MeFOSAA N-ethyl-d5 -perfluoro-1-octane-sulfonamidoacetic acid d5-N-EtFOSAA 13 Sodium 1H,1H,2H,2H-perfluoro-[1,2- C2 ] hexane sulfonate M2-4:2FTS Sodium 1H,1H,2H,2H-perfluoro-[1,2- 13C2 ] octane sulfonate M2-6:2FTS 13 Sodium 1H,1H,2H,2H-perfluoro-[1,2- C2 ] decane sulfonate M2-8:2FTS 1H,1H,2H,2H-Perfluoroctanyl acrylate M6:2FTA 1H,1H,2H,2H-Perfluorodecyl acrylate M8:2FTA 1H,1H,2H,2H-Perfluorododecyl acrylate M10:2FTA 2H-perfluoro-[1,2- 13C2 ]-2-octenoic acid M6:2FTUA 13 2H-perfluoro-[1,2- C2 ]-1-decenoic acid M8:2FTUA 2,3,3,3-tetrafluoro-2-(1,1,2,2,3,3,3-heptafluoropropoxy)- 13 MHFPO-DA C3 -propanoic acid Sodium bis(1H,1H,2H,2H-[1,2- 13C2 ] perfluorodecyl)phosphate M4 8:2 diPAP 66 3.2 Experimental section 3.2.1 Chemicals and reagents The PFAS standards, surrogates, internal standards, and other chemicals used for this study are described in Tables 3.1, 3.2, and 3.3. For a description of PFAS classes acronyms, see Table A1 in the Appendix. Table 3.2. Target PFAS, acronym, and surrogate standards for analysis by LC-QToF. Chemical Name Acronym Surrogate Standard Perfluoro-n-butanoic acid PFBA 1 M3PFBA Perfluoro-n-pentanoic acid PFPeA M3PFPeA Perfluoro-n-hexanoic acid PFHxA MPFHxA Perfluoro-n-heptanoic acid PFHpA M4PFHpA Perfluoro-n-octanoic acid PFOA M4PFOA Perfluoro-n-nonanoic acid PFNA MPFNA Perfluoro-n-decanoic acid PFDA MPFDA Perfluoro-n-undecanoic acid PFUdA MPFUdA Perfluoro-n-dodecanoic acid PFDoA MPFDoA Perfluoro-n-tridecanoic acid PFTrDA MPFDoA Perfluoro-n-tetradecanoic acid PFTeDA M2PFTeDA Perfluoro-n-hexadecanoic acid PFHxDA M2PFHxDA Perfluoropropane sulfonate PFPrS M3PFBS Perfluorobutane sulfonate PFBS M3PFBS Perfluoropentane sulfonate PFPeS M3PFBS Perfluorohexane sulfonate PFHxS MPFHxS Perfluoroheptane sulfonate PFHpS MPFHxS Perfluorooctane sulfonate PFOS MPFOS Perfluorononane sulfonate PFNS M3PFOS Perfluorodecane sulfonate PFDS M3PFOS Perfluorododecane sulfonate PFDoS M3PFOS 8-chloro-perfluorooctane sulfonate Cl-PFOS M3PFOS Perfluoroethylcyclohexane sulfonate PFEtCHxS M3PFHxS Perfluorobutane sulfonamide FBSA M8FOSA-I Perfluorohexane sulfonamide FHxSA M8FOSA-I Perfluorooctane sulfonamide FOSA M8FOSA-I N-methylperfluoro-1-octane sulfonamide MeFOSA d-N-MeFOSA-M N-ethylperfluoro-1-octane sulfonamide EtFOSA d-N-EtFOSA-M Perfluorooctane sulfonamido acetic acid FOSAA d3-N-MeFOSAA 67 Table 3.2. (cont’d) Chemical Name Acronym Surrogate Standard N-methylperfluorooctane sulfonamido acetic acid MeFOSAA d3-N-MeFOSAA N-ethylperfluorooctane sulfonamido acetic acid EtFOSAA d5-N-EtFOSAA 4:2 fluorotelomer sulfonate 4:2 FTS M2-4:2FTS 6:2 fluorotelomer sulfonate 6:2 FTS M2-6:2FTS 8:2 fluorotelomer sulfonate 8:2 FTS M2-8:2FTS 10:2 fluorotelomer sulfonate 10:2 FTS M2-8:2FTS 3:3 fluorotelomer carboxylic acid 3:3 FTCA M6:2FTA 5:3 fluorotelomer carboxylic acid 5:3 FTCA M6:2FTA 7:3 fluorotelomer carboxylic acid 7:3 FTCA M8:2FTA 6:2 fluorotelomer carboxylic acid 6:2 FTCA M6:2FTA 8:2 fluorotelomer carboxylic acid 8:2 FTCA M8:2FTA 10:2 fluorotelomer carboxylic acid 10:2 FTCA M10:2FTA 2H-Perfluoro-2-octenoic acid (6:2) 6:2 UFTCA M6:2FTUA 2H-Perfluoro-2-decenoic acid (8:2) 8:2 UFTCA M8:2FTUA dodecafluoro-3H-4,8-dioxanonanoate ADONA MPFNA 9-chlorohexadecafluoro-3-oxanonane-1-sulfonate 9Cl-PF3ONS MPFOS 11-chloroeicosafluoro-3-oxaundecane-1-sulfonate 11l-PF3OUdS MPFOS 2,3,3,3-tetrafluoro-2-(1,1,2,2,3,3,3- HFPO-DA MHFPO-DA heptafluoro propoxy)-propanoic acid bis(1H,1H,2H,2H-perfluorooctyl) phosphate 6:2diPAP M4 8:2 diPAP bis(1H,1H,2H,2H-perfluorodecyl) phosphate 8:2diPAP M4 8:2 diPAP bis-[2-(N-ethylperfluorooctane-1- diSAmPAP M4 8:2 diPAP sulfonamide) ethyl] phosphate 1 MRM transitions of 213 to 169 and 217 to 172 were used for quantification of PFBA and MPFBA, respectively, to reduce background. 3.2.2 Sample collection A leachate (labeled L1) was collected in September 2020 from a landfill in Michigan USA. The landfill accepts 850 tons of waste dialy and generates a leachate volume of 30000 gal/day. The landfill accepts municipal solid waste, commercial waste, construction & demolition waste, and non-hazardous industrial waste, with municipal solid waste as the most contributing waste fraction. The sample (20 L) was collected during the unloading process from a landfill leachates collection vehicle to a wastewater treatment plant. During sampling, nitrile-gloves were used to avoid 68 Table 3.3. Suspect PFAS detected in leachate L1. Chemical name Acronym Perfluorocyclohexan carboxylic acid PFCHxCA Perfluoromethyl cyclopentane carboxylic acid PFMeCPeCA Perfluropropyl cyclopentane sulfonate PFPrCPeS Perfluoromethyl cyclopentane sulfonate PFMeCPeS Unsaturated perflurohexane sulfonate UPFHxS Unsaturated perfluooctane sulfonate UPFOS Pentafluorosulfide-perfluoropentanoic acid F5S-PFPeA Perfluropropyl sulfonamide FPrSA Methyl perfluorobutane sulfonamide MeFBSA Methyl perfluorometane sulfonamido acetic acid MeFMeSAA Methyl perfluoropropane sulfonamido acetic acid MeFPrSAA Methyl perfluoropentane sulfonamido acetic acid MeFPeSAA Perfluorobutane sulfonamido acetic acid FBSAA Perfluoropentane sulfonamido acetic acid FPeSAA Ethyl perfluoropropane sulfonamido acetic acid EtFPrSAA Ethyl perfluorobutane sulfonamido acetic acid EtFBSAA Ethyl perfluorohexane sulfonamido acetic acid EtFHxSAA 12:2 fluorotelomer sulfonate 12:2 FTS 14:2 fluorotelomer sulfonate 14:2 FTS 4:3 fluorotelomer carboxylic acid 4:3 FTCA 2:2 fluorotelomer thia propanoic acid 2:2 FTThPrA 6:2 fluorotelomer sulfonyl propanoic acid 6:2 FTSO2 PrA Perfluoropentane sulfinate PFPeSi Perfluorohexane sulfinate PFHxSi Hydrido-perfluorobutane sulfonate H-PFBS Hydrido-perfluorohexane sulfonate H-PFHxS sample cross-contamination. The sample was collected using high-density polyethylene (HDPE) containers that were pre-rinsed with methanol. Following sample collection, the leachate was secured in coolers and shipped to the Fraunhofer USA Center Midwest, Division for Coatings and Diamond Technologies at Michigan State University. The leachate was stored at 3 °C upon receipt and experiments were performed immediately afterwards. 69 Power supply Current Voltage Electrochemical cell Peristaltic pump Reservoir tank Figure 3.1. Electrochemical oxidation setup used for the electrochemical treatment of leachate L1 3.2.3 Electrochemical oxidation setup Experiments were performed at laboratory scale with an electrochemical cell using niobium- supported polycrystalline BDD anodes and cathodes with high-boron doping level (Condias, Ger- many). The characterization of the cell is described in Appendix 3A. The cell utilized a series of rectangular parallel electrodes (3 anodes, 2 cathodes) of identical dimensions (200 × 26 × 2 mm), separated by 3 mm channels. The electrodes were connected in parallel. The total surface area of the anodes was 213.2 cm2 . Note that a parallel-plate cell was used for this chapter instead in the flow-through cell (used in previous chapter) to treat leachates. The change in configuration was attributed to the larger area ( 6 ×) available in the parallel-place cell that was expected to be beneficial for a faster PFAS transformation in leachates. All experiments were performed under galvanostatic conditions, in batch mode (2 L), and with single samples. A PVC tank was used as the reservoir/feed tank. The solution to treat (L1) was recirculated through the electrochemical cell with a flow rate of 2 L/min using a peristaltic pump. No pretreatment (e.g., filtration) was applied before the electrochemical process. Current densities of 10 and 50 mA/cm2 were used. Power was supplied by a BK−Precision 9202 (60 V × 15 A) power supply. The experimental setup is shown in Figure 3.1. Pure water was recirculated through the system for one hour without the application of 70 current to guarantee the absence of PFAS in all the components of the electrochemical reactor setup. The final effluent was sent for PFAS analysis. A control experiment to test for PFAS losses not attributable to electrochemical treatment was performed by recirculating L1 without the application of current. Gas sampling was not considered in this work. During the electrochemical experiments, L1 was monitored as a function of time. Typically, 10 mL of sample were collected every 2 h, transferred to polypropylene tubes, and stored in the refrigerator at −20◦ C until delivered for PFAS analysis. Additional parameters including pH, conductivity, and total organic carbon (TOC) were also monitored over time. The conductivity of L1 was 20.7 ± 0.2 mS/cm. Therefore, the addition of electrolyte was not required. 3.2.4 Analytical methods The TOC was determined using USEPA approved HACHTM standard methods. The pH and con- ductivity were measured with a SG23-B SevenGo DuoTM Series Portable Meter (Mettler Toledo). Addition of sodium Sample shaking upon chloride (2 g) and addition of 10% Addition of ethylene Titration of leachate surrogate PFAS trifluoroethanol (TFE) glycol (20 uL) to to pH 7-8 standards (10 uL) to 6 in ethyl acetate (EA) sampling vial mL aliquots (800 uL) Removal of top Addition of 10% TFE Reconstitution of Nitrogen blowdown organic layer (500 uL) in EA (500 uL) and extracted samples evaporation of from tube to a centrifugation. Repeat with DI water (50 uL) extracted samples sampling vial once. and methanol (70 uL) Addition of 10 uL of internal (M2PFOA LC-QToF analysis of and M8PFOS) samples standards to reconstituted samples Figure 3.2. Process diagram of the extraction method used for leachates 71 The temperature and flow rate were monitored using an in-house designed control system. Fluoride (F – ) and perchlorate (ClO4 – ) were analyzed via ion chromatography using EPA Methods 9056A and 314.0, respectively. The detection limit for F – quantification in leachates was 50 μg/L. Single replicates of each time point, with the exception of untreated L1 (t = 0), were used to generate the F – data. Standard error as a measure of the precision about the reported concentrations was calculated using replicate samples of untreated L1 (n = 3). For time points different than t = 0, propagated relative standard error was used as a measure of the uncertainty in the F – concentration. 3.2.4.1 PFAS quantification PFAS samples from preliminary experiments, control samples to test for the absence of PFAS in the electrochemical setup, and no-current controls to test for PFAS losses not attributable to electrochemical treatment were sent for PFAS analysis with a modified EPA 537 method to Eurofins TestAmerica (Sacramento, U.S.). The description of the method is found in Chapter 2. Control samples showed no PFAS present in the electrochemical setup and no-current controls showed negligible PFAS losses. Samples for target and suspect PFAS analysis were shipped to Oregon State University (OSU) for liquid chromatography quadrupole time-of-flight mass spectrometry (LC-QToF) analysis. The PFAS data from LC-QToF was produced from single replicates of each time point, with the exception of the untreated L1 (t = 0). Replicate samples of untreated L1 (n = 3) were used to calculate a standard error as a measure of precision about the reported concentrations. Propagated relative standard error was used as a measure of the uncertainty for time points different than t = 0. PFAS extraction and analysis with LC-QToF The PFAS analysis was performed using a sample extraction method adapted from Allred et al. [18] with some modifications. Figure 3.2 depicts a process diagram of the extraction method. Briefly, 6 mL of landfill leachate were placed into 15 mL polypropylene centrifuge tube and tritrated to pH 7-8 with 1 or 8 M of sodium hydroxide (NaOH) and/or 1 or 6 M of hydrochloric acid (HCl), depending on the buffer capacity of the sample. Sodium 72 chloride (2 g) was added to the samples and 10 μL of surrogate PFAS standards were spiked (see Table 3.1 for surrogate standards list). Samples were shaked for 30 s upon the addition of 800 μL of 10% trifluoroethanol in ethyl acetate. If there was no separation of layers, the sample was centrifuged for 2 min at 1625 rpm. The top organic layer from the tube (500 μL) was removed and transferred to an autosampler vial containing 20 μL of ethylene glycol. Next, the samples were centrifuged upon the addition of 500 uL of 10% trifluoroethanol in ethyl acetate (TFE). The last step was repeated once. The extracts were blown down to ethylene glycol under nitrogen, and reconstituted with 50 μL of deionized water and 70 μL of methanol (MeOH). Samples were transferred to an autosampler vial containing 10 μL of internal (M2PFOA and M8PFOS) standards. Chromatographic separation of PFAS was accomplished using an Agilent 1260 series LC fitted with a Zorbax NH2 and Sil guard columns, in-line with a Zorbax Eclipse Plus C18 analytical column (4.6 × 75 mm × 3.5 μm). The composition of the mobile phases were 3% methanol in water (A) and 10 mM ammonium acetate in methanol (B). All solvents were HPLC grade. A SCIEX X500R QToF-MS/MS system (Framingham, MA) was operated in negative electrospray ionization (ESI – ) mode. Data was collected using SWATH ®data-independent acquisition for both TOF-MS and MS/MS modes. PFBA and MPFBA were analyzed in MS/MS mode to reduce background. Precursor ion data (TQF-MS) were collected over a m/z range of 100 Daltons (Da; TOF start mass) to 1250 Da. The accumulation time was 200 ms and the ion spray voltage was -4500 V. Source and gas parameters included: a source temperature of 550 ◦ C, ion source gasses at 60 psi, curtain gas at 35 psi, and collision gas at 10 psi. The declustering potential was -20 V (with 0 V spread) and the collision energy was -5 V (with 0 V spread). Product ion scan (TOF-MS/MS) data were collected for a m/z range from 50 Da (TOF start mass) to 1200 Da. The accumulation time for each SWATH® window was 50 ms. Tables 3.2 and 3.3 shows a list of the target and suspect PFAS analyzed. 73 Table 3.4. Target PFAS, acronym, accuracy (% recovery), precision (% RSD), and limits of detection and quantification in landfill leachate by LC-QToF. The ‘*’ indicates that the surrogate standard was used in place of the target to estimate the LOD/LOQ. b ND indicates no surrogate available and target PFAS was in background leachate. Accuracy Precision LOD LOQ Chemical Name Acronym (% Recovery) (%RSD) (ng/L) (ng/L) Perfluoro-n-butanoic acid* PFBA* ND ND 9.7 29 Perfluoro-n-pentanoic acid* PFPeA* ND ND 16 48 Perfluoro-n-hexanoic acid* PFHxA* ND ND 8.4 25 Perfluoro-n-heptanoic acid* PFHpA* ND ND 3.5 10 Perfluoro-n-octanoic acid* PFOA* ND ND 11 32 Perfluoro-n-nonanoic acid* PFNA* 99 6.6 18 54 Perfluoro-n-decanoic acid PFDA 109 5.1 17 52 Perfluoro-n-undecanoic acid PFUdA 106 2.4 5.3 16 Perfluoro-n-dodecanoic acid PFDoA 108 4.6 6.3 19 Perfluoro-n-tridecanoic acid PFTrDA 90 12 10 31 Perfluoro-n-tetradecanoic acid PFTeDA 110 18 3.9 12 Perfluoro-n-hexadecanoic acid PFHxDA 113 17 4.3 13 Perfluoropropane sulfonate PFPrSa 138 6.2 3 10 Perfluorobutane sulfonate* PFBS* 88 14 5.1 15 Perfluoropentane sulfonate PFPeSa 85 14 7.6 25 Perfluorohexane sulfonate* PFHxS* 69 23 5.2 15 Perfluoroheptane sulfonate PFHpSa 165 4.2 3 10 Perfluorooctane sulfonate* PFOS* 92 9.9 15 45 Perfluorononane sulfonate PFNS 110 27 15 46 Perfluorodecane sulfonate PFDS 100 7.1 6.6 20 Perfluorododecane sulfonate PFDoS 75 8.8 8.7 26 8-chloro-perfluorooctane Cl-PFOS 100 20 12 35 sulfonate Perfluoroethylcyclohexane PFEtCHxS 89 8.0 5.1 15 sulfonate Perfluorobutane sulfonamide FBSA 28 22 11 34 Perfluorohexane sulfonamide FHxSA 34 9.4 12 35 Perfluorooctane sulfonamide FOSA 90 5.2 4.5 14 N-methylperfluoro-1-octane MeFOSA 111 11 9.7 29 sulfonamide N-ethylperfluoro-1-octane EtFOSA 104 5.4 7.2 22 sulfonamide Perfluorooctane sulfonamido FOSAA 95 11 7.2 22 acetic acid 74 Table 3.4. (cont’d) Accuracy Precision LOD LOQ Chemical Name Acronym (% Recovery) (%RSD) (ng/L) (ng/L) N-methylperfluorooctane MeFOSAA* 85 6.3 6.8 21 sulfonamido acetic acid N-ethylperfluorooctane EtFOSAA* 121 30 9.3 28 sulfonamido acetic acid 4:2 fluorotelomer sulfonate 4:2 FTS* 228 15 7.9 24 6:2 fluorotelomer sulfonate 6:2 FTS* 77 13 2.3 6.8 8:2 fluorotelomer sulfonate 8:2 FTS* 98 2.7 14 42 10:2 fluorotelomer sulfonate 10:2 FTS 102 22 7.0 21 3:3 fluorotelomer carboxylic acid 3:3 FTCAa 68 105 7.6 25 5:3 fluorotelomer carboxylic acid 5:3 FTCAa ND ND 3 10 7:3 fluorotelomer carboxylic acid 7:3 FTCAa ND ND 3 10 6:2 fluorotelomer carboxylic acid 6:2 FTCA* ND ND 16 47 8:2 fluorotelomer carboxylic acid 8:2 FTCA* 86 79 17 51 10:2 fluorotelomer carboxylic acid 10:2 FTCA 103 11 29 86 2H-Perfluoro-2-octenoic acid (6:2) 6:2 UFTCAa 38 3.2 3 10 2H-Perfluoro-2-decenoic acid (8:2) 8:2 UFTCA 89 7.3 8.1 24 dodecafluoro-3H-4,8- ADONA 63 13 12 35 dioxanonanoate 9-chlorohexadecafluoro- 9Cl-PF3ONS 120 32 7.9 24 3-oxanonane-1-sulfonate 11-chloroeicosafluoro- 11l-PF3OUdS 97 4.3 2.6 7.7 3-oxaundecane-1-sulfonate 2,3,3,3-tetrafluoro-2-(1,1,2,2,3,3,3- HFPO-DA 101 18 8.6 26 heptafluoro propoxy)-propanoic acid bis(1H,1H,2H,2H-perfluorooctyl) 6:2diPAP 785 43 9.5 29 phosphate bis(1H,1H,2H,2H-perfluorodecyl) 8:2diPAP 156 51 10 31 phosphate bis-[2-(N-ethylperfluorooctane-1- diSAmPAP 365 46 6.9 21 sulfonamide)ethyl]phosphate a The LOD/LOQ values were determined based on the original calibration curve and the quality control standards used throughout the analytical run. * The surrogate standard was used in place of the target to estimate the LOD/LOQ. Since the target PFAS was in the water, sample was used to determine the LOD/LOQ. b ND indicates no surrogate available and target PFAS was in background leachate. ND due to high background relative to overspike. 75 The method for target PFAS quantification with LC-QToF allowed for a recovery range of 70-130%. The solvent blank was LOQ are represented. PFAS with concentrations 90% of the PFSA molar composition. Concentrations of perfluorooctanoic acid (PFOA) and perfluorooctanesulfonic acid (PFOS), cur- 79 Table 3.6. Characterization for leachate L1a Parameter L1 Conductivity (mS/cm) 20.7 ± 0.2 pH 8.1 ± 0.1 Chemical oxygen demand (mg/L) 4250 ± 290 TOC (mg/L) 2300 ± 12 Fluoride (mg/L) 2.7 ± 0.4 Chloride (mg/L) 1800 ± 16 Total PFAAs (ng/L) 15500 ± 560 Total identified PFAS (ng/L) 51400 ± 3300 PFAA precursors from TOP assay (ng/L) 39000 ± 1700 a Standard error (SE) based on measurement of n =3 influent replicates. rently regulated by the Environmental Protection Agency (EPA), were 28 and 6 times above the EPA health advisory level (HAL) of 70 ng/L, respectively. PFAA precursors The PFAA precursors detected were divided in two groups: electrochemical fluorination (ECF) and fluorotelomer (FT) derived PFAS. ECF derived PFAS The ECF derived PFAS comprised 7% of the molar composition of L1. The ECF precursors included N-alkyl sulfonamido-acetic acids (N-alkyl FASAAs), perfluoroalkyl sulfonamides (FASAs) and other PFAS that do not fit in traditional categories and are referred from (a) Perfluoro cyclohexane (b) Perfluoromethyl carboxylic acid (PFCHxCA) cyclopentane carboxylic acid (PFMeCPeCA) Figure 3.4. Possible isomers of PFCHxA detected in leachate L1. 80 (a) Unsaturated (b) Perfluoro cyclohexane (c) Perfluoromethyl perfluorohexanoic acid sulfonate (PFCHxS) cyclopentane (UPFHxS) sulfonate (PFMeCPeS) Figure 3.5. Possible isomers of UPFHxS detected in leachate L1. now as "Other PFAS". The N-alkyl FASAAs present in L1 comprised MeFASAA (C1, C3, C5, C8), EtFASAA (C3, C4, C6, C8), and FASAA (C4, C5, C8), where Cn indicates the number of carbons with at least 1 F. The FASAAs likely arise from the transformation under methanogenic conditions of methyl- and ethyl perfluoroalkyl sulfonamidoethanols, which are used as intermediates in the synthesis of other PFAS and fluoropolymers [4, 13]. The identified FASAs included FPrSA (C3), FBSA (C4), FHxSA (C6), and FOSA (C8). The compound MeFBSA (C4) was the only MeFASA detected. Previous studies in leachates have only detected C6 and C8 FASAs [12]. The FASAs with C4 and C6 have been reported in groundwater and biota in far lower concentrations than in leachates [40–42]. The other PFAS derived from ECF comprised multiple cyclic PFAS (PFEtCHxS, PFPrCPeS, PFCHxA, and PFCHxS), unsaturated PFAS (UPFAS); pentafluorosulfide-perfluoroheptanoic acid (F5S-PFPeA), hydrido-perfluoroalkane sulfonate (H-PFAS); and perfluoroalkane sulfinates (PFASi) (see Table 3.5). Some of the cyclic PFAS presented isomers (Figures 3.4 and 3.5). The compound PFEtCHxS (traditionally used as an erosion inhibitor) is the only cyclic PFAS that has been previ- ously detected in landfill leachates [19]. The ECF derivatives F5S-PFPeA, H-PFAS, UPFAS, and PFASi have been found in commercial products and AFFF-impacted groundwater [43]. 81 FT based PFAS The FT based PFAS comprised 60% of of the molar composition of L1. The classes detected included FTCAs, FTSs, and other PFAS. The other PFAS with FT chemistry included the classes n:2 FDThP (C2) and n:2 FTSO2 PA (C6). Within the FTCA class, 3:3 FTCA, 4:3 FTCA, 5:3 FTCA, 6:2 FTCA, and 6:2 UFTCA, 7:3 FTCA, and 8:2 FTCA were detected and their concentrations in L1 were higher than previously reported values in landfill leachates [15, 18]. Compounds from the FTCA class have been reported to appear as biodegradation products of FTOH under anaerobic conditions, consistent with the operation of lined landfills, which are common in the US [44]. Within the FTSs class, 6:2 FTS, 8:2 FTS, 12:2 FTS, and 14:2 FTS were identified. Multiple FTSs have been reported to be released from consumer products, surfactants, detergents, and food packaging containing fluorotelomer-based substances [45, 46]. The other PFAS detected were 2:2 fluorotelomer thio propanoic acid (2:2 FTThPrA); and 6:2 fluorotelomer sulfonyl propanoic acid (6:2 FTSO2 PrA). The PFAS 2:2 FTThPrA and 6:2 FTSO2 PrA are FTCA and FTS derivatives, respectively, that have only been detected in groundwater [43, 47]. (a) (b) 5 60,000 ∑PFCAs or PFSAs after TOP assay Concentration (ng/L) PFDA 4 PFNA PFHpA PFOA 40,000 PFPeA 3 PFHxA PFBA ∑PFCAs or PFSAs before TOP assay 2 20,000 1 0 Pre-oxidation Post-oxidation Difference 0 PFCAs PFSAs Figure 3.6. (a) Concentration of PFCAS or PFSAs after TOP assay with respect to their concentra- tion at t = 0. (b) Concentration of PFCAs that resulted from the TOP assay. Results were obtained from n = 3 replicates. 82 Table 3.7. Influent PFAS concentrations in L1 (ng/L and nM) ± standard error and summed masses ± propagated standard error Method PFAS ng/L nmol/L PFCAs 10800 ± 500 37 ± 2 PFSAs 4700 ± 200 15 ± 0.5 LC-MS/MS PFAAs 15500 ± 600 52 ± 2 PFAA precursors 36000 ± 3200 105 ± 10 Total PFAS 51500 ± 3300 157 ± 10 PFCAs Before TOP Assay 12600 ± 200 42 ± 1 TOP Assay PFCAs After TOP Assay 50100 ± 1900 217 ± 8 Net production PFCAS from 39000 ± 1700 178 ± 7 PFAA precursors Total PFAS 54500 ± 1800 229 ± 7 (PFAAs + PFAA precursors TOP Assay) To the best of our knowledge, FBSA and most of the suspect PFAS are reported here for the first time in a landfill leachate, expanding the range of known PFAS present in this matrix. 3.3.1.2 PFAS contribution from TOP Assay The mean concentration of PFCAs after the TOP assay increased by 4-fold with respect to their pre-oxidation concentration (Figure 3.6a), revealing the presence of precursors. Results from the difference between post- and pre-oxidation revealed a PFAA-precursors concentration of 39000 ± 1700 ng/L. The predominant PFCA generated was PFBA (Figure 3.6b). Interestingly, the quantification of the concentration of PFAA-precursors with LC-QToF (105 nmol/L; Table 3.7) accounted for 59% of the concentration of precursors determined with TOP assay (178 nmol/L; Table 3.7). Assuming that the precursors identified with LC-QToF were oxidized in the post-oxidation step of the TOP assay, 41% of the total concentration of precursors with TOP assay (73 nmol/L) are unknown. The latter value of unknown precursors was added to the total mass balance of PFAS in untreated L1, bringing the initial concentration of total PFAS in L1 to 229 ± 7 nmol/L (see 3.7 for complete complete mass balance information of the influent PFAS concentrations in L1). 83 It is important to mention that the TOP assay method does not account for any precursors that are not oxidizable or that oxidize to substances other than C4−C10 PFCAs. In addition, due to the large amounts of salts produced during the method used [36], the quantification of PFCAs with chain lengths shorter than C4, which in a previous study showed to be representative in the quantification of precursors [48], is not included. Volatile precursors may not be captured by the TOP assay [49]. Identifying additional precursors not on the suspect lists by non-target liquid chromatography-high resolution mass spectrometry (LC-HRMS) was beyond the scope of this study. 3.3.2 PFAS transformations during electrochemical oxidation of L1 The electrochemical transformation of the PFAS identified in L1 was investigated. The leachate was electrochemically treated with a current density of 10 mA/cm2 that led to a voltage of 4.8 V. The pH remained circumneutral. Figure 3.7a shows the evolution of the detected PFAS classes over time. The total PFAS concentration (molar basis) increased 1.8-fold (420 ± 26 nmol/L) after 8 h of treatment with respect to the initial total PFAS concentration (229 ± 7 nmol/L). The latter reveals that the concentration of precursors are underestimated by both the LC-QToF analyses and the TOP assay. Non target analysis is needed to identify the unknown PFAS that contribute to this increase, but it is beyond the scope of this study. The transformation of unidentified precursors led to increasing trends of PFCAs, PFSAs, FASAs, and n:2 UFTCAs (Figure 3.8). The most notable increase was for PFCAs (Figure 3.7b) and FASAs (Figure 3.8e) by 8-and 71-fold, respectively, by the end of the treatment. The PFBA concentration comprised >80% of the PFCAs at t = 8 h. The increase of PFCAs in environmental matrices has shown to be a result of the transformation of PFAA precursor compounds [32, 36]. The PFSAs concentration increased by 50%, with PFBS as the most abundant (Figure 3.7c). Clearly, the TOP assay was not able to quantify the missing concentration of PFCA precursors that were electrochemically oxidized to form their transformation products detected at t = 8 h. The concentration of each target and suspect PFAS over time is shown in Figure 3.8. Note that PFAS with concentrations that decreased to 80% (Figures 3.8g, 3.8h). At the end of the electrochemical treatment with 10 mA/cm2 , 28% of the PFAS composition were PFAA precursors. Therefore, the current density was increased to 50 mA/cm2 to find the point where all PFAA precursors are transformed to PFAAs. The evolution of PFAS classes with 50 mA/cm2 (Figure 3.9a) significantly changed relative to 10 mA/cm2 . The most noticeable difference was the faster transformation of FASAs and n:3 85 (a) (b) MeFASAA EtFASAA n=1 n=3 n=5 n=4 1.0 1.0 PFAS concentration (C/C0) PFAS concentration (C/C0) 0.8 0.8 0.6 0.6 0.4 0.4 0.2 0.2 0 0 0 2 4 6 8 0 2 4 6 8 Time (h) Time (h) (c) (d) FASAA MeFASA n=4 n=4 1.0 20 PFAS concentration (C/C0) PFAS concentration (C/C0) 0.8 15 0.6 10 0.4 5 0.2 0 0 0 2 4 6 8 0 2 4 6 8 Time (h) Time (h) (e) FASA n=3 n=4 n=6 80 PFAS concentration (C/C0) 60 40 20 0 0 2 4 6 8 Time (h) Figure 3.8. Concentration of PFAS over time with respect to their initial concentration (t = 0 h) during the electrochemical treatment of leachate L1 with 10 mA/cm2 . n = number of carbons with at least 1 F – . The chemical structures are the general structures of each class. 86 Figure 3.8. (cont’d) (f) (g) n:2 FTS n:3 FTCA n=6 n=12 n=14 n=3 n=5 n=7 6 1.0 PFAS concentration (C/C0) PFAS concentration (C/C0) 0.8 4 0.6 0.4 2 0.2 0 0 0 2 4 6 8 0 2 4 6 8 Time (h) Time (h) (h) (i) n:2 FTCA n:2 UFTCA n=6 n=8 n=6 1.0 6 PFAS concentration (C/C0) PFAS concentration (C/C0) 0.8 5 0.6 4 0.4 3 0.2 2 0 1 0 2 4 6 8 0 2 4 6 8 Time (h) Time (h) (j) (k) UPFAS H-PFAS n=3 n=5 n=2 1.0 1.0 PFAS concentration (C/C0) PFAS concentration (C/C0) 0.8 0.8 0.6 0.6 0.4 0.4 0.2 0.2 0 0 0 2 4 6 8 0 2 4 6 8 Time (h) Time (h) 87 Figure 3.8. (cont’d) (l) (m) Cyclic PFAS n:2 diPAP PFCHxA or PFMeCPeCA n=6 15 PFAS concentration (C/C0) PFPrCPeS PFAS concentration (C/C0) 1.0 0.8 10 0.6 0.4 5 0.2 0 0 0 8 16 24 0 2 4 6 8 Time (h) Time (h) (n) (o) n:2 FDThP PFCAs n=2 n=3 n=5 n=7 1.0 PFAS concentration (C/C0) n=4 n=6 PFAS concentration (C/C0) 30 0.8 0.6 20 0.4 10 0.2 0 0 0 2 4 6 8 0 2 4 6 8 Time (h) Time (h) (p) PFSAs n=3 n=5 n=8 n=4 n=6 PFAS concentration (C/C0) 3 2 1 0 0 2 4 6 8 Time (h) 88 FTCAs. At the end of treatment, <0.4% of PFAA precursors (traces of FASAs, n:2 FTS, and H-PFAS) were part of the molar composition of the treated L1 that was dominated by PFCAs (92%) and PFSAs (7.6%). The compounds PFCHxCA, FBSA, and MeFBSA showed transient increases in concentration during treatment. However, the three compounds were degraded in >99% by the end of the treatment. The rest of PFAS compounds (excluding PFAAs) decreased over time and their final concentrations were 90% of organic co-contaminants present in L1. Although increasing the treatment time by a factor of 4 led to a decrease in the concentration of PFCAs, particularly the PFBA generated from precursors transformation, the 99 energy consumption associated with the longer treatment time also increased 4 times and could compromise the practicality of electrochemical treatment. Therefore, the extent to which PFBA should be degraded has to be considered for practical purposes. The ClO4 – generated corresponded to 67 and 165 mg/L after 8 h of electrochemical treatment with 10 and 50 mA/cm2 and increased to 1300 mg/L after 32 h of treatment with 50 mA/cm2 . Although the generation of ClO4 – was minimized with a low current densities (e.g., 10 mA/cm2 ), additional alternatives that prevent its generation should be considered. 3.3.6 Conclusions This work identified multiple PFAS in landfill leachates for the first time, highlighted the ability of electrochemical oxidation to treat PFAS, and showed evidence of known electrochemical degra- dation pathways. The results collectively suggested that precursors present in L1 corresponded to >75% of the concentration profile. The target FBSA and most of the suspect compounds were Table 3.8. Percentage removal of PFAAs in L1 after electrochemical treatment with multiple current densities. Negative values represent increase in concentration 10 mA/cm2 50 mA/cm2 50 mA/cm2 PFAAs 8h 8h 32 h PFBA -2100 -1800 26 PFPeA -220 -190 59 PFHxA -23 29 84 PFHpA -4 46 72 PFOA 24 81 86 PFNA > 99.9 > 99.9 > 99.9 PFDA > 99.9 > 99.9 > 99.9 PFPrS 12 1 -110 PFBS -68 -47 -140 PFPeS > 99.9 > 99.9 -48 PFHxS -41 -15 -280 PFHpS > 99.9 > 99.9 > 99.9 PFOS 0 62 -120 100 identified for the first time in leachates. The electrochemical treatment of L1 led to the generation of multiple transformation products that allowed for the identification of electrochemical degra- dation pathways. In brief, sulfonamide-based precursors and fluorotelomer-based precursors were electrochemically transformed into perfluoroalkyl carboxylic acids (PFCAs) during treatment of the leachate, consistent with previous literature. In addition, results after the electrochemical treatment showed the dominance in the con- centration of short-chain PFAS, in particular PFBA and PFBS. The extent to which PFBA and PFBS should be degraded determined the necessary treatment time and energy consumption of the electrochemical process. This important consideration should not be neglected in feasibility studies. Further, given the complexity of leachates and the much higher concentrations of a myriad of other compounds with respect to PFAS, pre-treatment technologies are necessary prior to the electrochemical treatment of PFAS in landfill leachates to increase the energy efficiency and reduce the treatment time of the electrochemical process. Although this work provided a preamble of the implications of electrochemical oxidation of PFAS in leachates, additional research is required to selectively oxidize PFAS and improve the feasibility of electrochemical oxidation for complex matrices. 101 APPENDICES 102 APPENDIX 3A CHARACTERIZATION OF BDD ANODES 103 Figure 3A.1. Scanning electron microscopy (SEM) image of the BDD surface The BDD anodes were characterized using scanning electron microscopy (SEM) and electro- chemical methods for capacitance and electrode kinetics. The results for the SEM characterization are depicted in Figure 3A.1 and show a high surface area film with conglomerated multi size grains and no cracks. The electrochemical capacitance was determined using cyclic voltammetry (Figure 3A.2). A constant area of the electrode was exposed to a 1 M potassium chloride (KCl) solution. The current that resulted from the application of a potential sweep from -0.5 to 0.5 V was measured for multiple scan rates. The capacitance was determined at 0 V and corresponded to 120 ± 2 uF/cm2 . 10 8 Current (μA) 6 Trial 1 4 Trial 2 Trial 3 0.1 0.2 0.3 0.4 0.5 Scan rate (V/s) Figure 3A.2. Current vs. scan rate plot to determine the capacitance of BDD 104 4×10−4 2×10−4 Current (A) 0 0.1 V/s −2×10−4 0.2 V/s 0.3 V/s 0.4 V/s −4×10−4 0.5 V/s −0.5 0 0.5 Scan rate (V/s) Figure 3A.3. Cyclic voltammogram of ferrocyanide on the BDD surface The kinetic constants of the electrodes were determined using potassium ferri/ferro cyanide in a 1M KCl solution (Figure 3A.3). The current that resulted from the application of a potential sweep from -0.4 to 0.8 V was measured for multiple scan rates. The Nicholson and Shain method was –1 used to determine the rate constant that corresponded to 1.65 (± 0.03) × 10 cm/s. 105 APPENDIX 3B PFAS CHARACTERIZATION OF L1 106 Table 3B.1. PFAS characterization of L1. Analytes include target and suspect PFAS. "n" represents the number of C with at least 1 F. Concentration values correspond to the average ± standard error. Target (T) or Class name Class structure n PFAS compound Concentration (ng/L) suspect (S) 3 PFBA T 2300 ± 170 4 PFPeA T 1700 ± 200 5 PFHxA T 3700 ± 400 Perflurocarboxylic acids 6 PFHpA T 1100 ± 33 (PFCAs) 7 PFOA T 1900 ± 200 8 PFNA T 60 ± 10 9 PFDA T 98%), perfluorooctanoic acid (PFOA, >98%), perfluorohexanesulfonic acid (PFHxS >98%), perfluorobutanoic acid (PFBA, >98%), potassium ferricyanide (K4 Fe(CN)6 ), potassium ferrocyanide (K3 Fe(CN)6 ), sodium carbonate (Na2 CO3 ), and sodium chloride (NaCl) were pur- chased from Sigma Aldrich, St. Louis, MO, USA. A synthetic still bottoms solution and a real still bottoms solution were used in the experiments. The real solution corresponded to AFFF contami- nated groundwater that was treated with IX resins. The composition of the solutions is described in Tables 4.1 and 4.2. The real still bottoms solution was provided by Emerging Compounds Treatment Technologies (ECT2) and shipped to the Fraunhofer USA Center Midwest at Michigan State University. Samples were stored at 4 °C upon receipt. 4.2.2 Electrochemical Oxidation Setup The laboratory and semi-pilot scale experiments were performed within two separate in-house build systems comprised of an electrochemical cell equipped with boron-doped diamond (BDD) 121 Table 4.1. Characterization of the synthetic still bottoms solution used for the electrochemical treatment of PFAAs in both laboratory and semi-pilot scales Compound Value pH 7.7 Conductivity (mS/cm) 110 PFBA (mg/L) 74 PFOA (mg/L) 86 PFHxS (mg/L) 87 PFOS (mg/L) 81 Chemguard C301 MS AFFF (% ) 0.1 Chloride (mg/L) 41670 Methanol (mg/L) 10000 TOC (mg/L) 2400 Table 4.2. Characterization of the real still bottoms solution used for the electrochemical treatment of PFAAs in laboratory scale Compound Value pH 9.7 Conductivity (mS/cm) 81.3 4:2 FTS (mg/L) 1.4 6:2 FTS (mg/L) 35.0 8:2 FTS (mg/L) 0.4 PFBA (mg/L) 95.8 PFPeS (mg/L) 0.3 PFHxS (mg/L) 98.0 PFHpA (mg/L) 0.3 PFHpS (mg/L) 0.3 PFOA (mg/L) 88.4 PFOS (mg/L) 59.3 Chloride (mg/L) 41,000 Methanol (mg/L) 28,000 TOC (mg/L ) 14,050 rectangular-plate electrodes (Condias, Germany), power supply, peristaltic pump, reservoir tank, pH, temperature and flow rate sensors. Table 4.3 and Figure 4.1 show details of the experimental 122 Table 4.3. Specifications of the electrochemical setup at laboratory scale and semi-pilot scale Parameter Laboratory Scale Semi-Pilot Scale Number of cathodes 2 5 Number of anodes 3 6 Inter-electrode gap (mm) 3 2 Electrode width (mm) 26 82 Anode area (cm2 ) 200 1400 Solution volume (L) 2 14 Flow rate (L/min) 2 6 setup for both scales. The semi-pilot scale setup was built by increasing the exposed anodic surface area of the laboratory scale by a factor of 7 and maintaining a constant area-to-volume ratio (A/V) for the treated solution. A flow rate of 6 L/min for the semi-pilot-scale setup was estimated by calculating the equiv- alent Reynolds number (Re) when compared to the laboratory scale setup. The Re number was determined using Equation (4.1) that considers the linear velocity and equivalent diameter of a Figure 4.1. Experimental setup for the electrochemical oxidation of PFAAs from IX still bottoms at the (a) laboratory and (b) semi-pilot scales. 123 parallel-plate cell [26]: 2·𝑄 𝑅𝑒 = (4.1) 𝑣 · (𝑊 + 𝑆) where 𝑄 is the flow rate (m3 /s), 𝑣 is the kinematic viscosity (m2 /s), and 𝑊 and 𝑆 are the width of a rectangular plate and the inter-electrode gap. 4.2.3 Electrochemical Experiments All experiments were performed in duplicate, batch mode, and under galvanostatic conditions. For the laboratory scale experiments, different current densities (10, 25, and 50 mA/cm2 ) were evaluated to determine the optimum current density to treat a synthetic still bottoms solution. For the semi-pilot-scale experiments, only the optimum current density found with the laboratory scale setup was used. Control experiments, without the application of current were also performed in duplicate. Experiments were typically performed for 8 h and samples were collected over time. Typically, 10 mL of sample was collected at each time point, transferred to polypropylene tubes, and stored in the refrigerator at 4 ◦ C until delivered for PFAS analysis. The conductivity of all solutions used was sufficiently high and the addition of electrolyte was not necessary. 4.2.4 Analytical Methods During the electrochemical experiments pH, temperature, conductivity, flow rate, voltage, fluoride (F – ), total organic carbon (TOC), perchlorate (ClO4 – ), and PFAS were monitored over time. TOC was determined using USEPA approved HACHTM standard methods. F – was analyzed via ion chromatography using EPA Method 9056A, and ClO4 – was analyzed via ion chromatography using EPA Method 314.0. The pH and conductivity were measured with an SG23-B SevenGo DuoTM Series Portable Meter (Mettler Toledo). Temperature and flow rate were monitored using in-house designed control systems. PFAS analysis was performed following a modified EPA 537 method by Trident Labs, Inc (Holland, MI, USA). Briefly, water samples and quality control (QC) samples were spiked with internal standards. A solid phase extraction (SPE) proceure was performed using Waters Oasis 124 Table 4.4. Calibration standards used for PFAS detection Analyte Description MRL* Units 4:2 fluorotelomer sulfonate (4:2 FTS) 2.0 ng/L 6:2 fluorotelomer sulfonate (6:2 FTS) 20.0 ng/L 8:2 fluorotelomer sulfonate (8:2 FTS) 2.0 ng/L N-ethylperfluorooctanesulfonamidoacetic acid (N-EtFOSAA) 10.0 ng/L N-methylperfluorooctanesulfonamidoacetic acid (N-MeFOSAA) 10.0 ng/L perfluorooctane sulfonamide (FOSA) 10.0 ng/L Perfluorobutanesulfonic acid (PFBS) 2.0 ng/L Perfluorobutanoic acid (PFBA) 2.0 ng/L Perfluorodecanesulfonic acid (PFDS) 2.0 ng/L Perfluorodecanoic acid (PFDA) 2.0 ng/L Perfluorododecanoic acid (PFDoA) 2.0 ng/L Perfluoroheptanesulfonic Acid (PFHpS) 2.0 ng/L Perfluoroheptanoic acid (PFHpA) 2.0 ng/L Perfluorohexanesulfonic acid (PFHxS) 2.0 ng/L Perfluorohexanoic acid (PFHxA) 2.0 ng/L Perfluorononanesulfonic acid (PFNS) 2.0 ng/L Perfluorononanoic acid (PFNA) 2.0 ng/L Perfluorooctanesulfonic acid (PFOS) 2.0 ng/L Perfluorooctanoic acid (PFOA) 2.0 ng/L Perfluoropentanesulfonic acid (PFPeS) 2.0 ng/L Perfluoropentanoic acid (PFPeA) 2.0 ng/L Perfluorotetradecanoic acid (PFTeDA) 2.0 ng/L Perfluorotridecanoic acid (PFTrDA) 2.0 ng/L Perfluoroundecanoic acid (PFUdA) 2.0 ng/L 4,8-dioxa-3H-perfluorononanoate (ADONA) 2.0 ng/L Hexafluoropropylene oxide dimer acid (HFPO-DA) 2.0 ng/L * MRL = Minimum reporting limit. WAX cartridges. A mixture of ammonium hydroxide/methanol was used to elute PFAS from the sorbent into a collection vial. The extracts were concentrated to dryness using a nitrogen evaporator and then reconstituted in 1 mL of methanol. Samples were injected and ran on an Agilent LC-MS/MS system fixed with a C18 column to separate out various PFAS and a C18 delay 125 column. The MS used an ion funnel in the negative ion mode to analyze the PFAS compounds of interest. Data analysis was performed using the Agilent QQQ Quantitative Analysis software to compare the retention time, mass spectra, ion ratio, etc., of the samples with the internal standards and calibration standards. The accepted recovery limits for quantification ranged between 50 and 150% . The calibration standards used for PFAS quantification are shown in Table 4.4. The PFAS precursors 4:2 fluorotelomer sulfonate (4:2 FTS), 6:2 fluorotelomer sulfonate (6:2 FTS), 8:2 fluorotelomer sulfonate (8:2 FTS), N-ethyl perfluorooctane sulfonamido acetic acid (NEtFOSAA), and N-methyl perfluorooctane sulfonamido acetic acid (NMeFOSAA) were below detection levels (<2000 ng/L) for all the synthetic still bottoms due to the dilution factor used in this work (10,000 ×), which was necessary to achieve concentrations within the linear dynamic range and quantify PFAS. 4.3 Results and Discussion 4.3.1 Laboratory Scale Evaluation 4.3.1.1 General Observations During the electrochemical treatment of the synthetic still bottoms solution in a laboratory scale setup, the applied voltages for the current densities evaluated ranged from 4 to 8 V. The pH of the solution was 7.7 ± 0.1 . After 8 h of treatment, the pH decreased by 15% with 10 and 25 mA/cm2 , and increased by 5% with 50 mA/cm2 . The TOC removal was 19, 27, and 67% after 8 h of electrochemical treatment with 10, 25, and 50 mA/cm2 . The TOC evolution over time is depicted in Figure 4.2. No decrease in PFAAs concentrations was observed in the control (no-current) experiments, indicating that adsorption of PFAAs by the system components was not significant. However, a layer of foam was formed during all the electrochemical experiments due to the electrochemical generation of hydrogen and oxygen at the electrodes [27]. The layer of foam substantially decreased in thickness after 4 h and a small but persistent layer remained throughout the rest of the experimental time in all experiments. This will be addressed in following sections. 126 100 10 mA/cm2 25 mA/cm2 80 50 mA/cm2 TOC removal (%) 60 40 20 0 2 4 6 8 Time (h) Figure 4.2. TOC removal over time during the electrochemical treatment of a synthetic still bottoms solution. The applied current densities were: 10 mA/cm2 , 25 mA/cm2 , and 50 mA/cm2 . 4.3.1.2 Influence of Current Density on PFAAs Removal The release of CF2 moieties during the electrochemical oxidation of PFAAs leads to the generation of F – , which increases over time during the PFAAs degradation process. Figure 4.3b depicts the F – generation over time with multiple current densities. The pseudo-first-order fluoride generation rate constant and r2 values are shown in Table 4.5. The influence of the current density on the electrochemical oxidation of PFAAs in a synthetic still bottoms solution was studied at the laboratory scale. Figure 4.3a shows the decrease in PFAAs concentration over time with the application of multiple current densities. The decrease in concentration was proportional to the applied current Table 4.5. Values of fluoride pseudo-first order generation rate constants during the electrochemical treatment of PFAS in still bottoms Current density k Scale –2 –1 r2 (mA cm ) (s ) –5 lab 10 6.08× 10 0.9999 –5 lab 25 5.83× 10 0.9744 –5 lab 50 1.72× 10 0.8861 –5 lab (real still bottom) 50 4.89× 10 0.9919 –6 Semi-pilot 50 8.15× 10 0.8994 127 (a) (b) 1.0 35 Fluoride concentration (mg/L) 10 mA/cm2 25 mA/cm2 30 0.8 50 mA/cm2 25 0.6 Ct / C 0 20 0.4 15 10 0.2 5 0 0 0 2 4 6 8 0 2 4 6 8 Time (h) Time (h) Figure 4.3. (a) Decrease in total PFAAs concentration and (b) fluoride generation over time for the electrochemical oxidation of a synthetic spent regenerant solution with 10, 25, and 50 mA/cm2 . Error bars represent the standard deviation of replicates. density and led to a total PFAAs removal of 46, 75, and 99% with 10, 25, and 50 mA/cm2 after 8 h of treatment, respectively. The decrease in concentration of total PFAAs followed a pseudo-first-order degradation rate and the corresponding values for the surface area normalized rate constants (kSA ) are depicted in Table 4.6. Figure 4.4 shows the concentration values for individual PFAAs over time. In general, long- chain PFAAs decreased in concentration faster than short-chain PFAAs. The increase in current density allowed for a higher removal of short-chain PFAAs. PFBA presented the slowest removal rate of the PFAAs detected and although a current density of 10 mA/cm2 was not able remove it, Table 4.6. Values of surface area normalized pseudo-first order degradation rate constants for the electrochemical treatment of PFAS in from a synthetic still bottoms solution Current density ksa Scale –2 r2 (mA cm ) (m s – 1 ) –6 lab 10 2.02 × 10 0.9944 –6 lab 25 4.41 × 10 0.9846 –5 lab 50 1.37 × 10 0.9554 –6 lab (real still bottoms) 50 4.25 × 10 0.5343 –6 Semi-pilot 50 8.44 × 10 0.9317 128 (a) (b) 400 400 PFOA PFAS concentration (mg/L) PFAS concentration (mg/L) PFOS PFHpA 320 PFHxA 320 PFHxS PFPeA PFBA 240 PFBS 240 160 160 80 80 0 0 0 2 4 8 0 2 4 8 Time (h) Time (h) (c) 400 PFAS concentration (mg/L) 320 240 160 80 0 0 2 4 8 Time (h) Figure 4.4. Decrease in concentration of individual PFAAs over time during the electrochemical treatment of a synthetic still bottoms solution at the laboratory scale. The applied current densities were: (a) 10 mA/cm2 , (b) 25 mA/cm2 , and (c) 50 mA/cm2 . 50 mA/cm2 allowed for >95% removal of PFBA and >99% removal for the remaining PFAAs. In addition, during the electrochemical treatment with 25 and 50 mA/cm2 , the shorter-chain PFAAs— perfluoroheptanoic acid (PFHpA), perfluorohexanoic acid (PFHxA), and perfluoropentanoic acid (PFPeA)—presented transient increases in concentration. The latter results from the oxidation of the head group of longer-chain perfluorinated carboxylates and sulfonates that release CF2 moieties leading to shorter-chain PFAAs, which are consecutively oxidized under the same unzipping mechanism [22, 28]. Although the concentration of F – increased with the applied current density, the generation rate 129 15 2 10 mA/cm 2 25 mA/cm 2 50 mA/cm Defluorination (%) 10 5 0 2 4 8 Time (h) Figure 4.5. Defluorination percentage during the electrochemical treatment of a synthetic still bottoms solution with 10, 25 and 50 mA/cm2 . constant was inversely proportional to the applied current density (Table 4.5) and PFAAs removal. For instance, the F – generation rate was 3.5-fold slower with 50 mA/cm2 when compared to 10 mA/cm2 . In addition, the F – concentration values were used to quantify the defluorination percentage over time. The defluorination values are shown in Figure 4.5. The values were calculated using Equation (4.2): 𝐶𝐹 [𝑡] − 𝐶𝐹 [0] 𝐷𝑒 𝑓 𝑙𝑢𝑜𝑟𝑖𝑛𝑎𝑡𝑖𝑜𝑛(%) = Í (4.2) 𝑛 𝐹,𝑖 × (𝐶0 − 𝐶𝑡 )𝑡 where 𝐶𝐹 [𝑡] and 𝐶𝐹 [0] are the concentrations of F – (mM) at time t and 0, respectively; 𝐶0 and 𝐶𝑡 are the concentrations of PFAAs (mM) at time 0 and t, respectively; and 𝑛 is the number of fluorine atoms in each PFAAs molecule present in the treated solution [24]. Similarly to the trend observed with F – generation, the defluorination percentage increased with the applied current density. However, this trend was true only for the first 2 h of treatment. The defluorination percentage determined for higher treatment time points was independent of the applied current density and the average value for all the applied current densities was 10.2 ± 0.9% and 12.6 ± 0.6% for 4 and 8 h of electrochemical treatment, respectively. Nevertheless, the defluorination values with different current densities were statistically different (p < 0.05) for all treatment times. Low defluorination ratios for still bottoms electrochemical treatment were also observed by Wang et al. [24]. The decrease in the F – generation rate with higher current densities and the low defluorination 130 values attained during the electrochemical treatment can be attributed to multiple factors. One of them is the inhibition of defluorination due to the high concentration of brine that corresponded to 4% NaCl for the synthetic solutions. Schaefer et al. evaluated the impact of different brine solutions on the defluorination in the electrochemical treatment of PFAS and observed a lower F – release for high concentrations of NaCl when compared to other brine solutions [22]. Both chloride Cl – oxidation and PFAAs defluorination occurs through direct anodic oxidation [29, 30]. In addition, the defluorination of PFAAs is rate-limited by direct oxidation at the anode surface [22]. Therefore, the low defluorination rate of PFAAs is likely attributed to the competitive reaction for chloride oxidation that ultimately leads to ClO4 – generation, which was shown to be the primary Cl – transformation product [22]. Incomplete oxidation of PFAAs, evidenced by the generation of shorter-chain PFAAs (Figure 4.4), was also ascribed to the low defluorination percentages. Other factors including recombination of F – with additional constituents in the solution, generation of unknown byproducts (e.g., fluoroalkane), and possible calcium fluoride (CaF2 ) precipitation could be associated with the low defluorination values. However, further investigation is required. The discrepancies between the high removal percentage and low defluorination rates of PFAAs with 25 and 50 mA/cm2 could have arisen due to the fact that some PFAAs were partially removed due to their accumulation in the layer of foam that was generated during the electrochemical experiments. The high concentrations of PFAAs, together with the electrochemically generated hydrogen and oxygen, likely facilitated foam partitioning. Therefore, a percentage of the removal of the highly hydrophobic PFAAs could have been attributed to their accumulation in the foam. This hypothesis is discussed in Section 4.3.2. 4.3.1.3 Electrochemical Treatment of Real Still Bottoms The current density that allowed for the highest PFAAs removal in the synthetic still bottoms solution (50 mA/cm2 ) was used to treat a real still bottoms sample at the laboratory scale. The treatment time was increased to 24 h to guarantee removal of short-chain PFAAs, given their slower degradation kinetics [31]. Figure 4.6 depicts the concentration of individual PFAS over time. 131 Long-chain PFAAs, short-chain PFAAs, and PFAA-precursors were present in the sample. The PFAS characterization of the sample is depicted in Table 4.2. Removal efficiencies were higher for long-chain PFAAs than for short-chain PFAAs. After 24 h of treatment, the concentration of total PFAS was reduced by 93%. In particular, long-chain PFAAs were removed by 95% , short-chain PFAAs by 87%, and PFAA precursors by 99%. Transient increases were observed for perfluorobutanesulfonic acid (PFBS), perfluoropentanoic acid (PFPeA), pefluoropentanesulfonic acid (PFPeS), perflueorohexanoic acid (PFHxA), and perfluoroheptanoic acid (PFHpA), likely ascribed to the degradation of precursor compounds and longer-chain PFAAs [18, 32]. Moreover, the kSA for total PFAS degradation was determined and corresponded to 4.3 × 10 −6 m/s, 7-fold lower than the kSA obtained for the synthetic spent regenerant solution (1.4 × 10 −5 m/s) treated with the same current density. A plausible explanation for the slower kinetics for PFAS removal in the real still bottoms is the interference of the additional organic matter and co-contaminants present in the matrix. The presence of organic matter and co-contaminants inter- feres with the electrochemical degradation process of target contaminants, usually by competitive oxidation [23, 33, 34]. The slower removal of PFAS was in accordance with a slower TOC removal (a) (b) 100 30 3 PFAS concentration (mg/L) PFAS concentration (mg/L) PFAS concentration (mg/L) PFNA PFPeS PFOS PFPeA 80 PFOA PFBS 25 2 PFHpS PFBA PFHpA 8:2 FTS 1 PFHxS 6:2 FTS 20 60 PFHxA 4:2 FTS 0 15 0 4 8 12 Time (h) 16 20 24 40 10 20 5 0 0 0 4 8 12 16 20 24 0 4 8 12 16 20 24 Time (h) Time (h) Figure 4.6. (a) Concentration of individual PFAS during the electrochemical treatment of a real still bottoms sample. The applied current density was 50 mA/cm2 . (b) Concentration of individual PFAS with concentrations lower than 30 mg/L. Inset depicts the evolution of PFAS with concentrations lower than 3 mg/L. 132 (shown in Fig 4.2), which was reduced by 18.5% after 8 h of treatment of the real still bottoms solution compared to 67% in the synthetic solution. Lastly, unlike the synthetic still bottoms, the real solution presented high concentration of PFAA-precursors that had to be oxidized together with PFAAs, adding more organic content to the solution. 4.3.2 Semi-Pilot-Scale Evaluation The laboratory-scale setup was scaled up by a factor of 7, while maintaining the A/V ratio used in the laboratory scale constant. The A/V ratio had a value of 10 m – 1 (0.02 m2 /0.002 m3 for the laboratory scale and 0.14 m2 /0.014 m3 for the semi-pilot scale). Prior to the evaluation of PFAAs removal, a mass transfer study was performed to determine the average mass-transfer coefficient (km ) in both setups (km,lab for the laboratory scale and km,sp for the semi-pilot scale). The values of km were determined with Equation (4.3), using the limiting-current technique—the procedure is described elsewhere [26, 35]. 𝐼𝑙𝑖𝑚 𝑘𝑚 = (4.3) 𝑛𝐹 𝐴𝐶𝐵 where 𝐼𝑙𝑖𝑚 is the limiting current (A), 𝑛 is the number of e – exchanged, 𝐹 is Faraday’s constant (96,485 C/mol), 𝐴 is the anodic area (m2 ), and 𝐶𝐵 is the concentration in the bulk (mol/m3 ). Constant concentrations of potassium of 0.05 M K4 Fe(CN)6 and 0.1 M K3 Fe(CN)6 were used for all the experiments. The concentration of K3 Fe(CN)6 was in excess to ensure the limiting current was at the anode. For the corresponding flow rates (2 L/min at the laboratory scale and 6 L/min at the semi-pilot scale) that provided an equivalent Re number for both setups ( 2300), km,lab was 7.0 × 10 −6 m/s and km,sp was 9.0 × 10 −6 m/s, giving a km,lab /km,sp ratio of 0.8. The value of km depends on the cell geometry and increases with a lower inter-electrode gap [26]. Therefore, the smaller inter-electrode distance of the semi-pilot scale (2 mm, compared to 3 mm at the laboratory scale) led to an enhancement of km at the semi-pilot scale. An enhancement in kSA for PFAAs degradation was also expected at the semi-pilot scale. Consecutively, the electrochemical treatment of PFAAs in a synthetic still bottoms solution was assessed at the semi-pilot scale and the results were compared with those obtained at the laboratory 133 scale. The voltage that resulted from the galvanostatic process was lower at the semi-pilot scale (5.7 V at the semi-pilot scale vs. 5.9 V at the laboratory scale), attributed to the smaller inter-electrode distance, as previously stated. Figure 4.7 shows the decrease in concentration of total PFAAs from the synthetic still bottoms treated with 50 mA/cm2 in both scales. The total PFAAs removal after 8 h of treatment was 94% in the semi-pilot-scale setup. The percentages of individual PFAAs remaining in solution after treatment with respect to their initial concentrations were 19% of PFBA, 3% of PFHxS, and <2% of PFOA and PFOS. Similar to the laboratory scale experiments, a layer of foam was observed during the electrochemical treatment of PFAAs in the semi-pilot-scale setup. Therefore, a fraction of PFAAs removal, in particular the highly hydrophobic PFAAs, was likely attributed to their partitioning into the foam. To verify this, the foam generated during the experimental time (8 h) was collected separately and sent for PFAS analysis. Results showed that the mass percentage of individual PFAAs partitioned into the foam with respect to the initial concentration of PFAAs in the solution corresponded to 61% of PFOS, 17% of PFOA, 8% of PFBA, and 2% of PFHxS. Likewise, a previous study showed that at least 80% of the PFOS-associated fluorine partitioned into the foam [22]. 1.0 0 Lab-scale Semi-pilot scale −1 0.8 ln (C / C0 ) −2 0.6 Ct / C 0 −3 0.4 −4 0 2 4 6 8 Time (h) 0.2 0 0 2 4 6 8 Time (h) Figure 4.7. Decrease in total PFAAs concentration during the electrochemical treatment of a synthetic still bottoms solution with 50 mA/cm2 in laboratory and semi-pilot scale systems. Inset shows the pseudo-first-order removal rate for PFAAs for both system scales. 134 (a) (b) (b) Fraction of Molar F in PFCAs Fraction of Molar F in PFSAs PFOA PFOS 1.0 PFHpA 1.0 PFHxS PFHxA PFPeS PFPeA PFBS PFBA 0.8 0.8 0.6 0.6 0.4 0.4 0.2 0.2 0 0 0 2 4 8 0 2 4 8 Time (h) Time (h) (c) (d) Fraction of Molar F in PFCAs Fraction of Molar F in PFSAs PFOA PFOS 1.0 PFHpA 1.0 PFHxS PFHxA PFPeS PFPeA PFBS PFBA 0.8 0.8 0.6 0.6 0.4 0.4 0.2 0.2 0 0 0 2 4 8 0 2 4 8 Time (h) Time (h) Figure 4.8. Fraction of molar F relative to 𝑡 = 0 in PFCAs and PFSAs during the electrochemical oxidation of a synthetic still bottoms solution with 50 mA/cm2 . (a, b) correspond to experimentation at the laboratory scale. (c,d) correspond to experimentation at the semi-pilot scale. The fraction of molar F in PFCAs and PFSAs (shown in Figure 4.8) was used to compare the evolution of individual PFAAs over time during the electrochemical treatment in both scales. In general, higher fractions of PFCAs, in particular PFHpA, PFHxA, and PFPeA, were generated at the laboratory scale, suggesting faster degradation kinetics at the laboratory scale and more foam partitioning at the semi-pilot scale. The values of kSA for total PFAAs removal were 1.4 × 10 −5 m/s and 8.4 × 10 −6 m/s for the laboratory and the semi-pilot scales, respectively, giving a kSA,lab / kSA,sp ratio of 1.6. Interestingly, opposite ratios showing kSA,lab > kSA,sp and km,lab < km,sp were obtained. 135 The lower value of kSA for PFAAs removal in the semi-pilot setup suggests that other factors besides fluid properties, hydrodynamics, and A/V ratio play a critical role in the treatment efficiency of PFAAs in IX still bottoms. These factors include gas evolution and current density distribution [36]. During the electrochemical oxidation of target compounds (e.g., PFAAs), only a fraction of the applied current density, equal to the limiting current, is used in the oxidation of the target compound [27]. The remaining fraction of current is used in side reactions including oxygen and hydrogen evolution [37]. The previous reactions generate substantial quantities of gas (𝑉𝑔𝑎𝑠 ) that are proportional to the applied current, according to Faraday’s first law of electrolysis (Equation (4.4)): 𝐼 𝑅𝑇𝑡 𝑉𝑔𝑎𝑠 = (4.4) 𝑛𝐹𝑃 where 𝑅 is the universal gas constant (8.314 J/mol – 1 K – 1 ), 𝐼 is the current applied (A), 𝑇 is the average working temperature (303 K), 𝑡 is the treatment time (s), 𝐹 is the Faraday’s constant (96,485 C/mol), 𝑃 is the atmospheric pressure (1 × 105 Pa), and 𝑛 is the number of e – exchanged (2 for H2 , and 4 for O2 ). A higher electrode area requires the application of a higher current to maintain a constant current density between both reactor scales, leading to the generation of a higher volume of gas in the semi-pilot-scale setup. For the corresponding currents of each setup (10.7 A at the laboratory scale and 70 A at the semi-pilot scale), the total volume of gas generated corresponds to 7.5 L/h and 49.3 L/h, approximately 7-fold more gas generation at the semi-pilot scale. Although a local increase in the mass transfer is expected if gas bubbles are generated [27], the inherent surface- active properties of PFAAs induce their movement towards the air-water interface of the bubbles [38], that travel to the interface of the solution (foam generation), where PFAAs are partitioned. In addition, local gas hold-up in the vicinity of the electrodes could have interfered with direct anodic oxidation of PFAAs in the liquid phase [37]. Therefore, the probability of PFAAs reaching the anode surface decreases, slowing down the oxidation process. Thus, a lower kSA is obtained. Finally, possible differences in current density distributions along the electrodes in each setup could have affected the mass transfer of the process [36, 39]. To maintain current similarity, it is rec- ommended to increase the number of smaller modules, rather than increase the electrode size [36]. 136 (a) (b) 20 20 Perchlorate concentration (mM) Perchlorate concentration (mM) Real sb laboratory scale 10 mA/cm2 25 mA/cm2 Synthetic sb laboratory scale 16 50 mA/cm2 16 Synthetic sb semi-pilot scale 12 12 8 8 4 4 0 0 0 2 4 6 8 0 2 4 6 8 Time (h) Time (h) Figure 4.9. Perchlorate generation during the electrochemical treatment of (a) synthetic still bottoms solution with 10, 25 and 50 mA/cm2 , (b) real still bottoms at the laboratory scale, synthetic still bottoms at the laboratory scale, and synthetic synthetic still bottoms in a semi-pilot scale with 50 mA/cm2 . 4.3.3 Perchlorate Formation during Electrochemical Treatment ClO4 – generation was quantified for all experiments and its evolution over time is shown in Figure 4.9. For the electrochemical treatment of synthetic solutions with multiple current densities, the zero-order generation rate of ClO4 – increased with the current density (Figure 4.9a and reached concentrations of 2.6, 10.0, and 16.1 mM after 8 h of treatment with 10, 25, and 50 mA/cm2 , respectively. ClO4 – concentrations at the end of the treatment time (8 h) accounted for 0.2, 0.8, and 1.4% of the initial Cl – concentration (1250 mM). The generation of ClO4 – was relatively low compared to the initial concentration of Cl – available for oxidation. Although the concentration of chlorate (ClO3 – ) was not quantified in this work, a recent study performed with still bottom solutions reported equimolar concentrations of ClO3 – and ClO4 – generated after 40 h of electrochemical treatment [24]. Even assuming equivalent concentrations of ClO3 – generated, the percentage of chlorinated byproducts remains low when compared to the initial concentrations of Cl – . These results suggest that additional species present in the solution may be competing for direct anodic oxidation or scavenging Cl – oxidation. Wang et al. showed that the presence of methanol (100−1000 mM) in still bottom solutions scavenges chlorine radical Cl• generation and significantly 137 reduces the formation of chlorinated byproducts [24]. The synthetic still bottoms solution of this work included a concentration of methanol of 312 mM, which likely contributed to the reduction of ClO4 – generation. Moreover, although having similar initial concentrations of Cl – , the generation rate of ClO4 – during the electrochemical treatment of the real still bottoms was 2-fold slower than with the synthetic solution (Figure 4.9b). The latter suggests that Cl• scavenging may be affected by additional constituents of the solution, besides methanol. However, this assumption requires further studies. Finally, under the same experimental conditions, the generation of ClO4 – at the semi-pilot scale was comparable to the results obtained at the laboratory scale (Figure 4.9b). The ClO4 – concentration after 8 h of electrochemical treatment was of 17.3 mM. The results suggest that ClO4 – generation with BDD electrodes solely depends on the applied current density, regardless of factors associated to scale performance differences. 4.3.4 Treatment Efficiency and Energy Consumption 0.0025 Lab scale Semi-pilot scale 0.0020 Current efficiency 0.0015 0.0010 0.0005 2 4 6 8 Time (h) Figure 4.10. Coulombic efficiency (CE) for fluoride generation during the electrochemical treat- ment of a synthetic still bottoms solution at the laboratory and semi-pilot scales. The applied current density was 50 mA/cm2 . The coulombic efficiency (CE) was used to quantify the current efficiency for PFAAs defluori- nation during the electrochemical treatment and it is defined in Equation (4.5): [40, 41] 138 𝐹𝑉 𝑒𝐶𝐹 𝐶𝐸 = (4.5) 𝐼𝑡 where 𝐹 is Faraday’s constant (96,485 C/mol), 𝑉 is the volume of solution treated (L), 𝑒 is the moles of e – needed per mole fluoride (1 electron per C-F bond [22]), 𝐶𝐹 is the fluoride concentration (mol/L), 𝐼 is the current (A), and 𝑡 is the treatment time (s). As shown in Figure 4.10, the CE decreases over time from 2.3 × 10 −3 at 2 h of treatment to 8.6 × 10 −4 at 8 h of treatment. A comparable but lower decreasing trend was observed at the semi-pilot scale, with 15 and 40% lower CE at 2 and 8 h of electrochemical treatment, respectively. The low and decreasing CE values, characteristic of mass-transfer limited electrochemical reactions with applied potentials above the water oxidation threshold, are attributed to competitive oxidation reactions from additional components of the solution (e.g., Cl – , additional TOC) and water electrolysis reactions [27]. Nevertheless, the reported CE values are 5-fold greater than the values reported for the electrochemical treatment of low concentrations of PFAS in groundwater [29], showing that the efficiency of the electrochemical treatment of PFAS increases with highly concentrated solutions, such as still bottoms from IX spent regenerant solutions. Finally, the electric energy per order (EEO ) was determined using Equation (4.6) as follows [42]: 𝑃𝑡 𝐸 𝐸𝑂 = (4.6) 𝑉 𝑙𝑜𝑔(𝐶/𝐶0 ) where 𝑃 is the power of the system (W), 𝑉 is the treatment volume (L), 𝑡 is the treatment time (h), and 𝐶0 and 𝐶 are the initial and final PFAAs concentration. The energy required for 90% PFAAs removal with a current density of 50 mA/cm2 was 173 and 194 Wh/L for the laboratory and semi-pilot scales, respectively. Although the smaller inter-electrode distance in the semi-pilot- scale system provided a lower voltage, the faster degradation kinetics in the laboratory scale setup compensated the energy losses that result from a wider electrode gap, leading to a lower energy consumption required for the same order of removal. The latter highlights the importance of a fast degradation rate in the electrochemical process that allows for energy optimization. Last, it is important to consider that the energy consumption for the electrochemical treatment of PFAS from still bottoms accounts for less than 0.01% of the total volume of water pre-treated with 139 IX resins [10]. Therefore, the calculated energy required for the electrochemical treatment of the total volume of pre-treated water with IX is 0.017 Wh/L at the laboratory scale and 0.019 Wh/L at the semi-pilot scale. This outcome illustrates the benefits of a combined tandem IX- electrochemical oxidation process that allows for >99.9% energy reduction for the combined IX/EO technologies when compared to electrochemical oxidation of PFAAs alone. 4.4 Conclusions This work focused on the evaluation of the electrochemical treatment of PFAAs from still bottoms at the laboratory and semi-pilot scales. Results at the laboratory scale showed >99% removal for total PFAAs, which included >95% removal for PFBA and >99% removal for PFOA, PFHxS, and PFOS, with 50 mA/cm2 after 8 h of electrochemical treatment. However, low defluorination values were reported. Competitive oxidation of Cl – and PFAAs foam partitioning were attributed as the main factors for low defluorination. Additionally, the electrochemical treatment of a real still bottoms solution allowed for 93% removal of PFAAs after 24 h of treatment. However, 3- fold slower degradation kinetics for PFAAs compared to the synthetic still bottoms solution were measured, likely due to the presence of additional co-contaminants in the matrix. The results from the semi-pilot scale presented slower degradation kinetics for total PFAAs removal with respect to the laboratory scale and allowed for 94% of total PFAAs removal after 8 h of treatment. Minimization of foaming and scaling up of smaller modules, rather than increasing the electrode size may help to improve the similarity between scales that provide an equivalent performance. The generation of ClO4 – was not affected by the scale of treatment and corresponded to <2% of the initial concentration of Cl – for both scales. Additionally, more than 99.9% of energy savings in electrochemical oxidation were estimated for the total volume of water treated with the IX, highlighting the benefits of combining tandem technologies. Moreover, the addition of an anti-foaming agent (e.g., alcohol) may be necessary to avoid PFAS foam partitioning and consequently improve PFAAs degradation kinetics. Increasing the concentration of alcohol in the still bottoms could eliminate foaming while simultaneously reduce 140 ClO4 – generation. If the previous approach is effective, the increase in alcohol concentration could be achieved by reducing the distillation time of the regenerant solutions, which likely will reduce the distillation cost, providing two benefits: cost reduction of the tandem treatment and enhanced efficiency of the EO process. 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Walsh, Industrial Electrochemistry, 2, second ed., Springer Netherlands, Dordrecht, 1993. [40] C. A. Martínez-Huitle, M. A. Rodrigo, I. Sirés, O. Scialdone, Single and Coupled Electro- chemical Processes and Reactors for the Abatement of Organic Water Pollutants: A Critical Review, Chem. Rev. 115 (2015). [41] C. Comninellis, G. Chen, Electrochemistry for the environment, 1 ed., Springer, New York, NY, 2010. [42] J. Radjenovic, B. I. Escher, K. Rabaey, Electrochemical degradation of the b-blocker meto- prolol by Ti/Ru0.7Ir0.3O2 and Ti/ SnO2-Sb electrodes, Water Res. 45 (2011). 146 CHAPTER 5 DIELECTROPHORESIS-ENHANCED ADSORPTION FOR THE REMOVAL OF PFOA FROM WATER 147 5.1 Introduction Per-and polyfluoroalkyl substances (PFAS) are a group of synthetic chemicals of growing concern due to their ubiquitous presence, persistence in the environment, and associated health effects [1, 2]. The continuous manufacture, use, and disposal of PFAS over the last eighty years have resulted in contamination of water sources with PFAS, with concentrations ranging from pg/L to μg/L [3, 4]. Perfluoroctanoic acid (PFOA) and perfluorooctane sulfonate (PFOS) are the two most studied PFAS due to their toxicity, recalcitrant nature, and prevalence in drinking water systems [5, 6]. Additionally, both substances are transformation products of multiple polyfluorinated precursors [7]. Human exposure to PFAS has been associated with multiple health effects (e.g., inmunotoxicity, neurotoxicity, testicular and kidney cancer) [3, 8]. Consequently, in 2016, the United States Environmental Protection Agency (USEPA) established a health advisory level (HAL) of 0.07 μg/L for the combined concentration of PFOA and PFOS in drinking water [9, 10]. Conventional water treatment processes (e.g., flocculation/sedimentation/filtration) and biolog- ical degradation are ineffective for removing PFAS [3, 11–13]. In addition, destructive technologies have only been proven to work at bench scale, which hitherto has limited their implementation in real applications at large-scale [14–16]. Given the high volumes to treat and low concentrations of PFAS, a continuous treatment that removes low levels of PFAS is required for drinking water. Adsorption technologies, attributed to their low energy cost and ease of implementation, have been adopted as an emerging solution for PFAS water contamination [12, 14]. Common adsorbents including granular activated carbon (GAC), powdered activated carbon (PAC), and carbon block technologies have been used in community and household water treatment [17]. The ability of activated carbon to remove long-chain PFAS has been widely documented [12, 18–20]. However, carbon adsorbents require long contact times and frequent replacement to guarantee the removal of PFAS [21]. In addition, most carbonaceous adsorbents exhibit a relatively low adsorbent-phase concentration for PFAS and are ineffective capturing short-chain PFAS[22, 23]. Processes involving uniform electric fields (e.g., electrosorption) have been shown effective in enhancing the adsorption rates and adsorbent-phase concentration of molecules [6, 24–26]. The enhancement arises from 148 the directional drift of charged molecules towards the oppositely charged adsorbent surface. In addition to the generation of a uniform electric field, the adsorption of contaminants could be further enhanced by generating a non-uniform electric field that creates a stronger directional drift as a result of a dielectrophoretic effect. The dielectrophoresis (DEP) principle has been used in adsorption [27, 28], electrocoagulation [29, 30], and filtration [31] processes to enhance the removal of contaminants in water. The DEP is a force that appears as a result of the application of a non- uniform electric field on the induced dipole moment of a particle, generating a translational motion of the particle towards a stronger or weaker electric field, where it is adsorbed (e.g., adsorption, filtration) or precipitates (e.g., electrocoagulation) [32]. Thus, the removal of contaminants is enhanced and the treatment time reduced. The scope of this work was to reduce the necessary contact times and enhance the removal of a commonly found PFAS molecule, PFOA, through a dielectrophoresis-enhanced adsorption process. The specific objectives include: i) evaluate and compare the PFOA removal that results from treating samples with adsorption only, a uniform electric field-enhanced adsorption, and a non-uniform electric field (dielectrophoresis)-enhanced adsorption process in batch mode; ii) assess the PFOA removal with a dielectrophoresis-enhanced adsorption process in continuous mode. This study demonstrates a highly effective electro-adsorption under a non-uniform electric field. 5.2 Materials and methods 5.2.1 Materials Perfluorooctanoic acid (PFOA, >97%) and ethyl cellulose were obtained from Sigma Aldrich. Powdered activated carbon (PAC)-YP-80F was purchased from Kuraray. A standard stock solution of PFOA was prepared by dissolving the solid standard in methanol. The volume ratio of methanol in aqueous solution of electrophoretic experiments was less than 0.1%. A synthetic PFOA solution with a concentration of 50 μg/L was prepared from the stock and was used for all the experiments. The composition of the solution only included PFOA and DI water and the pH of the solution was 6. 149 (a) (b) Figure 5.1. (a) Electrophoretic deposition (EPD) setup used for fabricating carbon-coated elec- trodes. (b) uncoated and coated carbon electrodes. 5.2.2 Fabrication of carbon-coated electrodes Carbon-coated electrodes (Figure 5.1) were fabricated using electrophoretic deposition (EPD) [33]. The PAC and ethyl cellulose were mixed in a weight ratio of 9:1 in isopropyl alcohol (IPA). The solution was ultrasonicated for 30 min to disperse the PAC and guarantee an homogeneous solution. A stainless-steel (SS) rod and a SS tube were used as anode and cathode, respectively. A rubber stopper on top and a PVC holder in the base of the cell allowed centered the anode. The cathode (SS tube) was filled with the carbon solution followed by immersion of the SS rod to be coated. A voltage of 100 V was applied to the SS rod for 5 min using a DC power supply (Kikusui PAN-600- 2A). Subsequently, the carbon coated SS rods were dried in air for 20 min followed by annealing at 120 ◦ C for 24 h to evaporate the residual IPA and improve the adhesion of the carbon to the SS rod. All the carbon-coated electrodes were fabricated under the same experimental conditions. 5.2.3 Characterization of electrodeposited electrodes The surface morphology of the PAC after the electrophoretic deposition was analyzed by scanning electron microscopy (SEM). The electrodeposited electrodes were cut into small pieces ( 2.0 cm × 150 length) and attached to a mount with carbon tape. The SEM images were obtained at an accelerating voltage of 12 kV, working distance of 12 mm and SS of 30. 5.2.4 Theory of dielectrophoresis drift of dipole Dielectrophoresis (DEP) is a well-established phenomenon that uses an electric field for separa- tion and manipulation of particles [32]. The DEP arises when a polarizable (dielectric) particle is subjected to a strong non-uniform electric field that can be generated through asymmetrical electrodes such as insulating hurdles, posts, and curvature configurations [31, 32]. The polarized particles (induced dipole moment) have Coulomb forces of different magnitudes acting on each of the particles sides, resulting in a net translational force, known as the dielectrophoretic force (𝐹𝐷𝐸 𝑃 ) [28], which is governed by Eq. 5.1: 𝐹𝐷𝐸 𝑃 = 2 · 𝜋 · 𝜖 𝑚 · 𝑟 3 · 𝑅𝑒[ 𝑓𝐶 𝑀 ] · ▽|E| 2 (5.1) Where 𝜀 𝑚 is the electrical permittivity of the suspending medium, 𝑟 is the radius of the dipole moment, 𝑅𝑒[ 𝑓𝐶 𝑀 ] is the real part of the Clausius-Mossotti factor, and ▽|𝐸 | 2 is the electric field gradient. The real part of the Clausius-Mossotti factor 𝑓𝐶 𝑀 can be calculated with Eq. 5.2: 𝜀 ∗𝑝 − 𝜀 ∗𝑚   𝑓𝐶 𝑀 = 𝑅𝑒 (5.2) 𝜀 ∗𝑝 + 2𝜀 ∗𝑚 Where 𝜀 ∗𝑝 and 𝜀 ∗𝑚 are the complex permittivities of the particle and the suspending medium, respectively, defined as 𝜀 ∗ = 𝜀 − 𝑗𝜎 𝜔 , where 𝜎 and 𝜔 are the conductivity and angular frequency √ of the applied electric field, respectively, and 𝑗 = −1 [34, 35]. A complex component of the permittivity is considered when working with AC power supplies. Only the real part is used for DC power supplies and Eq. 5.2 can be simplified in terms of the real conductivites as follows [36]:   𝜎𝑝 − 𝜎𝑚 𝑓𝐶 𝑀 = (5.3) 𝜎𝑝 + 2𝜎𝑚 The direction of movement attributed to the DEP is determined by the electrical permittivity of the fluid and particle [29]. If the particle has a greater electrical permittivity or conductivity 151 than the fluid (𝑅𝑒[ 𝑓𝐶 𝑀 ] > 0), it will experience a positive-DEP and move towards an area of higher electric field. If the particle has a smaller electrical permittivity or conductivity than the fluid (𝑅𝑒[ 𝑓𝐶 𝑀 ] < 0), it will experience a negative DEP and move towards a lower electric field [31, 36–38]. The DEP force affects all particles regardless of their electrical charge (e.g., negative charge, neutral) [29, 34]. In addition, the strength of the force depends strongly on the medium and particles’ electrical properties, shape and size of particles, as well as on the frequency of the applied electric field [34]. For two concentric cylindrical electrodes (configuration used in this work), the electric field can be calculated using Eq. 5.4: 𝐸 = −▽𝜑 (5.4) where, 𝜑 is the root mean square (rms) of the electrostatic potential. This term can be given by Laplace’s equation assuming that the medium is liquid and homogeneous 5.5: ▽2 𝜑 = 0 (5.5) It is important to mention that dielectrophoresis is not equal to electrophoresis (5.2). Although both describe the movement of particles under the influence of applied electric fields, the former arises due to the force that a neutral particle experiences in a non-uniform electric field and acts in the direction of the increasing field strength. The latter arises through the electrostatic attraction between a charged particle and an oppositely charged electrode and follows electric field lines [31]. Figure 5.2. Direction of particle translational motion under the influence of dielectrophoresis and electrophoresis [31]. 152 Figure 5.3. Configuration of the coaxial-electrode cell for water treatment in: (a) batch mode (b) continuous-flow mode. Both dielectrophoresis and electrophoresis can occur simultaneously if particles in the solution are charged [30]. 5.2.5 Dielectrophoresis-enhanced adsorption cell A tubular coaxial-electrode cell (CEC) (Figure 5.3) was used to generate a non-uniform electric field. The CEC consisted of a carbon-coated electrode (positive) at the center and a coaxial cylindrical electrode (negative). The CEC was used in batch (Figure 5.3b) and continuous flow (Figure 5.3b) modes. The only difference between experimental setups was the length of the CEC. A copper (Cu) tube (12.7 mm ID, 14 cm length-batch; 26.2 cm length-continuous flow) served as the outer negative electrode (cathode). The carbon-coated electrode (6.4 mm OD) was placed in the center along the Cu tube and served as the coaxial center positive electrode (anode). The interlectrode distance was 3.2 mm. A planar carbon-coated electrode with rectangular shape and an interelectrode distance of 3 mm was fabricated for comparison. During the experiments in batch mode, the CEC was filled with 14 mL of PFOA solution. Multiple voltages ranging from 0 to 50 V were applied for various periods of time. Sample aliquots were taken before (t = 0) and at the end of the experiments. During the experiments in continuous flow mode, 5000 mL of PFOA solution were pumped through the CEC with a constant flow rate of 50 mL/min, which corresponded to an hydraulic retention time (HRT) of 30 s. The effective 153 volume of the cell was 25 mL. Voltages ranging from 0 to 25 V were applied for a time period of 100 min. Sample aliquots were taken every 10 min. 5.2.6 Analytical methods The concentration of PFOA was quantified with ultra performance liquid chromatography (Waters Acquity I-class Plus UPLC) coupled with a Waters TQ-XS mass spectrometer. The PFOA separation was performed on an Acquity UPLC BEH C18 (2.1 × 50 mm, 1.7 µm) column. The mobile phase was 10 mM ammonium acetate in water (A) and acetonitrile (B) (A:B= 99:1) and had a flow rate of 5 uL/min. The TQ-XS mass spectrometer operated in negative ESI multiple reaction monitoring (MRM) mode. The parameters used for quantification of PFOA were: precursor mass m/z of 413, daughter ion mass m/z of 369, dwell time of 163 ms, cone voltage of 20 V, and collision energy of 10 V. The gradient elution was: 0 min (A=99%, B=1%), then ramp to (A=1%, B= 99%) at 4 min, next ramp to (A=99%, B=1%) at 5 min and kept until 7 min. An internal standard 13C8 PFOA was used for mass loss correction (precursor mass m/z of 421, daughter ion mass m/z of 376). The desolvation temperature, desolvation gas flow, and ion spray voltage were maintained at 400 ◦ C, 800 L/h, and 1000 V, respectively. The cone gas flow was 150 L/h and the nebuliser gas flow was 7 psi. 5.3 Results and discussion 5.3.1 Characterization of electrodes Figure 5.4 shows SEM images of the PAC after its electrophoretic deposition on the surface of the SS rod. The deposited PAC presented an agglomeration of smaller carbon particles on top of bigger particles, creating an heterogeneous surface. This agglomeration can be attributed to: i) the ability of smaller particles to stay in suspension for a longer time, and ii) the higher mobility of smaller particles under an electric field. In addition, a 100 × image (Figure 5.4a) revealed the presence of cracks in the PAC coating. Cracking of the coatings in electrophoretic deposition may arise from the difference between the substrates, affecting the corrosion resistance properties [39, 40]. 154 (a) (b) Figure 5.4. SEM image of carbon coated electrodes with (a) 100 × and (b) 2000 × Obtaining crack-free coatings requires an optimization of the process / materials. However, the optimization of the electrophoretic deposition was not part of the scope of this work. 5.3.2 Mechanisms of dielectrophoresis enhanced PFAS adsorption The performance of the CEC in batch mode (Figure 5.3a) was compared to a planar electrodes cell (PEC) to understand the mechanisms of a non-uniform electric field enhanced adsorption. The PEC cell consisted of two rectangular electrodes facing each other that generate a uniform electric field. A non-uniform electric field cannot be generated with the PEC configuration. Therefore, the generation of solely a uniform electric field on the removal of PFOA was expected with the PEC. The results (Figure 5.5) show that under the same applied voltage (25 V) and treatment time (2 min), the CEC led to a 9-fold increase in the adsorbent-phase concentration of PFOA (mg PFOA/g PAC) when compared to the PFOA removal with the PEC configuration. This outcome supports the contribution that the dielectrophoretic forces have on the adsorption of PFOA with the generation of a non-uniform electric field. The PFOA removal with adsorption only was also evaluated for each configuration. The adsorbent-phase concentration with adsorption only (0 V) was 3-fold higher with the CEC cell relative to the PEC cell. However, with the application of a potential, the generation of a uniform-electric field in the case of the PEC and a uniform and non-uniform electric field in the case of the CEC, led to a 11 and 3-fold increase in the adsorbent-phase concentration 155 6 0 V PEC 0 V CEC 5 25 V PEC qt (mgPFOA / gC) 25 V CEC 4 3 2 1 0 Figure 5.5. Adsorbent-phase concentration(mg PFOA/g PAC) that resulted from the application of a uniform and non-uniform electric field with 25 V of external voltage. PFOA0 = 50 μg/L. Error bars represent the standard deviation of n = 3 replicates of PFOA (mg PFOA/g PAC) in the PEC and CEC, respectively. The strength of a non-uniform electric field has shown to be higher when compared to a uniform electric field due to the existence of (FDEP ) [37, 41, 42], and it was reflected on the experimental results. Figure 5.6 shows the adsorption mechanism of PFOA molecules attributed to electric fields. PFOA dissociates into the perfluooctanoate anion and the hydrogen ion when dissolved in water over a wide range of pH conditions, attributed to its low dissociation constant (pKa =3.8) [43, 44]. Molecules with low pKa values of 4 or less, such as PFOA, exist in aqueous solutions at neutral pH (7) almost entirely as the dissociated acid [43]. Thus, PFOA is present in the anionic form (negatively charged) in environmental matrices. With the application of a positive voltage with + Anode (carbon coated electrode) PFOA molecules Cathode (copper electrode) - Figure 5.6. Principle behind the dielectrophoresis-enhanced adsorption of PFOA 156 the PEC, the negatively charged PFOA molecule is attracted to the oppositely charged electrode (anode, positive), resulting in the drift of the molecule towards the carbon coated anode. The latter force is known as electrophoresis and results from the electrostatic attraction of charged particles in a uniform electric field. The CEC creates a stronger electric field than the PEC under the same voltage and electrode distance. In addition to an uniform electric field, the configuration of the CEC allows for the generation of a non-uniform electric field that induces a dielectrophoresis force on water molecules attached to the PFOA anion. This force removes the attached water molecules and the PFOA anion-water cluster size is greatly reduced. A small cluster size enables a much faster drift of PFOA anions. On the other hand, a uniform electric field has no effect on the water dipoles and does not change the size of the PFOA anion-water clusters, leading to a slow drift of the cluster as a result of the large size and mass of the PFOA molecule. Although the electric field is a gradient, the strongest potential locates where the gradient is generated (center of the carbon-coated electrode). It has been shown that the diffusion energy barriers of molecules decrease significantly after applying an external electric field facilitating adsorption [45]. 5.3.3 Dielectrophoresis effect on the adsorption of PFOA in batch mode The asymmetrical configuration of the CEC design in batch mode (Figure 5.3a), allowed for the generation a non-uniform electric field near the center electrode through the application of a positive external voltage. The effect of the applied voltage (5, 25, and 50 V) and treatment time (2, 10, and 20 min) was evaluated during the dielectrophoresis-assisted adsorption of PFOA. For the evaluation of the applied voltage (Figure 5.7a), the PFOA removal percentage increased by 4, 7 and 8− fold with 5, 25, and 50 V, respectively, when compared to adsorption only (no voltage applied, 0 V), and corresponded to 12, 50, 86 and 95% removal with 0, 5, 25 and 50 V, respectively. Thus, the PFOA removal with 50 V was the greatest. Since the DEP force depends on the electric field intensity, ▽ |E| 2, the magnitude of the DEP force increases with the increase of voltage, leading to an enhanced adsorption of PFOA. Moreover, when the voltage is higher, the particle velocity 157 (a) (b) 100 100 80 80 PFOA removal (%) PFOA removal (%) 60 60 40 40 20 20 0 0 0 5 25 50 2 10 20 Voltage (V) Treatment time (min) Figure 5.7. PFOA removal percentage that resulted from the application of: (a) different voltages with a constant treatment time of 2 min and (b) different treatment times with a constant voltage of 25 V. Initial concentration of PFOA0 = 50 μg/L. Error bars correspond to the standard deviation of n = 3 replicates. increases, resulting in a faster adsorption process. An increase on the applied voltage, leads to an increase on the electric field gradient due to the to contribution of the DEP force [46]. The effect of the treatment time on the dielectrophoresis-enhanced adsorption of PFOA was also evaluated. A constant voltage (25 V) was used for all the experimental treatment times (2, 10, and 20 min). The results (Figure 5.7b) reveal that the increase of the treatment time to 10 min had no influence on the removal of PFOA with respect to the removal obtained at 2 h (86%). A further increase (20 min) only led to an additional 3% removal. These results suggest that the dielectrophoresis phenomenon is not time dependent and occurs instantaneously. A study conducted by Kadaksham et al. that used numerical simulations to study the behavior of particles under the influence of a non-uniform electric field determined that the maximum particle drift attributed to electric field related forces occurs in a matter of seconds [47]. Further, for the experimental setup of this work, the drift of PFOA molecules due to the effect of dielectrophoresis is followed by a slow diffusion in the adsorption process of PFOA molecules into the internal sites of the PAC. Slow diffusion has also been observed with similar processes such as electrosorption [48]. 158 5.3.4 Continuous flow-operation of the CEC A continuous-flow prototype of the CEC (Figure 5.3b) was built to increase the volume of water treated from 14 to 5000 mL. The performance of the CEC in continuous-flow mode was investigated using a fixed HRT of 30 s (flow rate of 50 mL/min) and multiple applied voltages. Figure 5.8a shows the PFOA concentration of the effluent that resulted from the treatment of a PFOA solution with an initial concentration of 50 μg/L with adsorption only (0 V), 5 and 25 V. In general, the increase of voltage allowed for a higher PFOA removal when compared to adsorption only and corresponded to 18, 22, and 44% removal for 1L of treated water. The increase of PFOA removal with higher voltages is attributed to the dielectrophoretic forces generated through the non-uniform electric field. The higher applied voltages are necessary to overcome the drag forces through the CEC, allowing the PFOA molecules to be removed from the stream of water [28]. However, the PFOA removal decreased with the treated volume in all the experimental conditions (higher PFOA concentrations in the effluent with the increase of treated volume). Although the PFOA concentration of the effluent decreased with the increase of voltage, after 3000 mL of solution (a) (b) 1.0 0.20 0.8 0.16 qtotal ( μg PFOA/mg C) CPFOA/CPFOA0 0.6 0.12 0.4 0.08 Control 0.2 5V 0.04 25 V 0 0 0 1000 2000 3000 4000 5000 0 5 25 Volume (mL) Voltage (V) Figure 5.8. (a) PFOA effluent concentration that resulted from the application of 0, 5 and 25 V and (b) total adsorbent-phase concentration (qtotal , μg PFOA/mg C) after the treatment of 5000 mL of a 50 μg/L PFOA solution with a non-uniform electric field-enhanced adsorption process with 0, 5 and 25 V of voltage using the CEC cell in continuous-flow mode. The applied flow rate was 50 mL/min. The error bars represent the standard deviation of n = 3 replicates. 159 treated, the application of 25 V had no effect on the enhancement of PFOA removal when compared to 5 V. At 5000 mL, the application of an electric field was irrelevant as the removal with or without an electric field was the same. The former may be a result of the saturation of the carbon-coated electrode, which is reaching its adsorption capacity due the acceleration of the adsorption process with higher voltages. The latter can be confirmed with a calculation of the adsorption capacity of the carbon coated electrode after treating 5000 mL of PFOA solution. The results (Figure 5.8b) shows that the adsorption capacity (qtotal , μg PFOA/mg C) increased 2.5 and 2.6-fold with 5 and 25 V when compared to adsorption only (0 V), highlighting the benefits of using a non-uniform electric field for the enhancement of PFOA adsorption in short contact times. It is important to note, however, that although the PFOA removal with the application of a non-uniform electric field in the continuous-flow cell increased by 1.2 and 2.4-fold with 5 and 25 V, respectively, with respect to adsorption only, the removal percentages in all cases were lower than with the batch cell (12, 50, and 86% for 0, 5, and 25 V). Among multiple factors, the lower removal percentages of PFOA are a result of the shorter contact time (HRT of 30 s, flow rate of 50 mL/min) applied in the continuous-flow cell when compared to the batch cell (HRT of 2 min). Low removal efficiencies at higher flow rates may result from the loss of adsorbed molecules caused by viscous drag forces. The dielectrophoretic separation is often described as a balance of competing forces that gives rise to a net force that dictates the movement of a molecule / particle [28]. If the forces that work to move the PFOA away from the regions of strong electric fields (e.g., viscous drag, diffusional and lift forces) are greater than the forces holding them in place in the carbon- coated electrode (e.g., DEP, dipole, gravitational), the PFOA molecules will not be adsorbed on the carbon-coated electrode. High flow rates increase the drag force, which consequently reduces the probabilities of PFOA molecules to be adsorbed by the carbon-coated electrode. Jun et al. showed an increase on the retention of bacteria on a dielectrophoretic cell when using low flow rates [28]. Therefore, higher removal percentages would require a much longer HRT (lower flow rate) to enhance the adsorption process [49]. 160 5.4 Conclusions The work presented herein utilized a CEC able to generate a non-uniform electric field that enhanced the adsorption of PFOA through the contribution of dielectrophoretic forces to the adsorption process. The application of an electric field increased by 4, 7, and 8-fold the PFOA removal with the application of 5, 25, and 50 V when compared to adsorption only. Moreover, a comparison between the generation of a uniform electric field and both a uniform and non-uniform electric field led to 11 and 3-fold increase in the adsorbent-phase concentration of PFOA (mg PFOA/g PAC) with respect to adsorption only. The evaluation of the CEC performance in continuous-flow mode for the removal of PFOA showed an increase of 1.2 and 4-fold in the PFOA removal with respect to adsorption only. However, the carbon-coated electrode reached faster its saturation point with the increase of voltage, which was reflected with a higher effluent concentration with the increase of volume treated. Lower flow rates are suggested for the improvement of the PFOA removal in continuous-flow mode. 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Bull. 59 (2014). 166 CHAPTER 6 CONCLUSIONS AND FUTURE DIRECTIONS 167 6.1 Summary and Conclusions The remediation of recalcitrant legacy contaminants such as per-and polyfluoroalkyl substances (PFAS) is challenging for drinking water and wastewater treatment facilities. Several treatment technologies to address the PFAS contamination in multiple matrices were evaluated in this thesis work. The first matrix addressed was landfill leachates. Landfill leachates are included in the number of impacted sources with PFAS contamination. Optimizing PFAS remediation in leachates is important as it could prevent PFAS from migrating to other water sources (e.g., groundwater) that expand PFAS contamination and expose humans to these contaminants. In addition, the use of destructive technologies, such as electrochemical oxidation, could be a potential solution to the increasing PFAS accumulation cycle. In the first part of this study (Chapter 2), the concentrations of PFAS in 6 different leachates from 3 landfills in Michigan were determined. The concentration of individual perfluoroalkyl acids (PFAAs) ranged from 10 2 to 10 4 ng/L. Perfluorocarboxylic acids (PFCAs) were in higher con- centrations than perflurorosulfonates (PFSAs). Perfluoroctanoic acid (PFOA) and perflurobutane sulfonate (PFBS) were identified as the PFAAs with the highest concentrations. Subsequently, a boron-doped diamond (BDD) flow-through cell was used to evaluate the electrochemical oxida- tion (EO) of PFAAs. The performance of the flow-through cell was assessed and compared with synthetic solutions for the oxidation of PFOA and PFOS. The electrochemical oxidation of various leachates with a current density of 150 mA/cm2 allowed for high removal efficiencies of long chain PFAAs but led to the generation of high concentrations of short-chain PFAAs, in particular, per- fluorobutanoic acid (PFBA), which generation was associated with the transformation of precursor compounds. This chapter provided information about the predominance and prevalence of PFAA precursors in landfill leachates, as well as the energy demands necessary to electrochemically treat PFAS in landfill leachates. In the second part of the study (Chapter 3), the transformation of PFAA-precursor compounds during the electrochemical oxidation of PFAS-impacted landfill leachates was investigated. A 168 leachate with high concentrations of precursor compounds was selected for this study. Target and suspect PFAS were identified in the leachate and their concentrations during electrochemical treatment were quantified over time. Liquid chromatography quadrupole time-of-flight mass spec- trometry (LC-QToF) measurements of the leachate allowed for the identification of 52 PFAS and 19 different classes. Multiple PFAS were reported for the first time in landfill leachates. The molar composition of the leachate was comprised of 33% PFAAs, 7% electrochemical fluorination (ECF) precursors, and 60% fluorotelomer (FT) precursors. Further analysis with total oxidizable precur- sor (TOP) assay revealed an additional concentration of precursors that was not identified with LC-QToF. The evaluation of the intermediate and final products generated during the electrochem- ical treatment showed evidence of known electrochemical degradation pathways. However, this is the first study to have more evidence for electrochemical pathways in landfill leachates. In brief, sulfonamide-based precursors and fluorotelomer-based precursors were electrochemically trans- formed into perfluoroalkyl carboxylic acids (PFCAs) during treatment of the leachate. This chapter provided evidence of multiple PFAS non-reported previously in landfill leachates. The knowledge generated in this chapter could benefit the scientific community in future research related to PFAS in landfill leachates. The second matrix addressed (Chapter 4) was PFAS-impacted groundwater. The groundwater of this study was pretreated with ion exchange (IX) resins that allowed to concentrate high levels of PFAS in a small volume. The waste of the IX still bottoms that included PFAS, traces of methanol and organic content was electrochemically treated at laboratory and semi-pilot scales. Synthetic and real solutions were included. Multiple current densities were evaluated at the laboratory scale and the optimum current density was used at the semi-pilot scale. The results at the laboratory scale showed >99% removal of total PFAAs. PFAAs treatment at the semi-pilot scale showed 0.8-times slower pseudo-first order degradation kinetics for total PFAAs removal compared to the laboratory scale, and allowed for >94% PFAAs removal. Defluorination values, perchlorate generation, coulombic efficiency, and energy consumption were also assessed for both scales. Overall, the results of this study highlighted the benefits of a tandem concentration/destruction 169 (IX/EO) treatment approach, specially regarding energy savings, and discussed the implications for the scalability of EO to treat high concentrations of PFAAs. This chapter provided an initial guideline for the scale-up of electrochemical processes targeted to PFAS in a treatment train approach and could some insights for scalability considerations. The third and last matrix addressed (Chapter 5) was drinking water. The technology assessed was dielectrophoresis-enhanced adsorption for the removal of low concentrations of PFOA from water (e.g., tap water). This study introduced a coaxial-electrode cell (CEC) that allowed for the generation of a non-uniform electric field to enhance the adsorption of PFOA. The enhancement of the process was attributed to the generation of dielectrohoretic forces. Experiments were performed in batch and continuous-flow mode. The dielectrophoretic-enhanced adsorption in batch mode led to 4, 7, and 8-fold increase in the removal of PFOA when compared to adsorption only. The performance of the CEC in continuous-flow mode allowed for an increase of 1.2 and 4-fold in the PFOA removal. Overall, the results evidenced the benefits of using a dielectrophoresis-enhanced adsorption process for the removal of PFOA from water. This chapter contributed with potential solutions to reduce the adsorption time of PFAS molecules, specifically PFOA. Overall, the four studies performed in this work contributed to the understanding of PFAS degradation in multiple matrices with electrochemical oxidation, and introduced an alternative process to enhance the widely used adsorption technology for PFAS removal, optimization that could solve some of the main challenges of the technology. Finally, the treatment implications of each matrix were discussed and provided a clear baseline for future research, development, and scale-up of treatment technologies for PFAS that could be eventually implemented. 6.2 Challenges encountered Some of the challenges encountered during the multiple studies for PFAS treatment included: • The electrochemical treatment of PFAS led to a low current efficiency (CE) and although the CE was improved in 5-fold with a pretreatment with IX, a further improvement is necessary. • Larger electrode areas were necessary to guarantee PFAA-precursors transformation in land- 170 fill leachates. • The degradation time for PFAS in complex matrices required of multiple hours of treatment and must be reduced if the process is considered for real application with larger volumes. • During the dielectrophoresis-enhanced process, the carbon electrodes used during the con- tinuous process reached their saturation point too early. 6.3 Future Directions The next recommended future directions for research on PFAS remediation technologies will help to overcome some of the challenges encountered in this work and advance the state of the art: Investigating pretreatment technologies that isolate PFAS from complex matrices Pretreat- ment technologies aiming to preconcentrate PFAAs present in complex matrices (e.g, leachates, wastewater) in a simpler and more selective matrix (e.g., PFAS and water only) may improve the technical and economical feasibility of destructive technologies, such as electrochemical oxidation. By reducing the treatment volume and eliminating some of the competitive species from matri- ces, the energy consumption and degradation efficiency of electrochemical technologies could be improved. Engineering and optimizing electrochemical cell designs to treat PFAS Optimizing the design of electrochemical cells to treat PFAS could reduce the cost of implementation of the technology in real applications. The enhancement of the mass transfer is imperative to further develop this technology. The ultimate goal of using electrochemical oxidation should be to apply it as a one-pass treatment technology. One of the options to improve the mass transfer might be to use an electrofiltration cell design where the solution flows through one electrofilter or multiple electrofilters in series during treatment that oxidizes contaminants in contact with the active area. With this configuration, more surface active area available could significantly improve the mass transfer of the process. 171 Further, it should be considered that the same electrochemical design cannot be applied for every environmental matrix. Although simpler designs like a parallel-plate cell might work for the electrochemical oxidation of PFAS with high concentrations, it might not necessarily work with trace levels of PFAS as the chances of PFAS reaching the surface area in a low concentration solution are lower. Thus, the process efficiency decreases. In general, every cell design should be used accordingly. Material optimization It is also necessary to modify/engineer/optimize the material used in electrochemical treatment of PFAS (e.g., BDD) to increase selectivity for PFAS oxidation and reduce the energy consumption of the process. As mentioned in the previous paragraph, the optimization of electrochemical cell designs is necessary but this can only be accomplished with a material that allows for this to happen. Unfortunately, the current fabrication methods of BDD only allow to produce flat sheets that led to designs such as the parallel-plate cell (Chapter 3) and flow-through cell (Chapter 4). A porous material able to significantly enhance the available surface area of the cell and the mass transfer of the contaminants passing through it is necessary for the development of this technology. In addition, future materials to be used for electrochemical oxidation, should consider the scaling up capabilities, that is to say, the ability to produce the material in bigger sizes. Study of catalysis alternatives for the electrochemical oxidation of PFAS The electrochemical oxidation of PFAS occurs at potentials above the water oxidation (e.g., 3.6 V). Thus, in addition to PFAS oxidation reactions, water oxidation reactions occur simultaneously, generating substantial quantities of hydrogen and oxygen in the form of gas, that leads to a low current efficiency. Thus, most of the energy consumed in the process is used in other reactions but PFAS oxidation. Alternatives (e.g., materials, solutions, coupling technologies or sources) aiming to catalyze the PFAS oxidation potential could significantly improve the selectivity and efficiency of the process. In addition, for the particular case of PFAS, given their surface-active properties, the generation of gas during the electrochemical process might slow down the degradation kinetics. The hydrogen 172 generated during the electrochemical process mobilizes a fraction of the most hydrophobic PFAS to the air-water interface. Therefore, gas generation must be minimized for PFAS treatment. Study of coupled hydrogen-generation/electrochemical oxidation process As shown in Chap- ter 4, the current efficiency of the electrochemical process of PFAS is low. The latter is a result of the high potential (above the water oxidation threshold) applied in the process that is necessary to oxidize PFAS and that leads to multiple reactions occurring at the same time, mainly water oxidation. During water oxidation, H2 is produced in large quantities. The H2 generated could be captured for its use in other processes. Therefore, the electrochemical oxidation of PFAS could be used as an opportunity to produce H2 in a multipurpose process. With this consideration, the low current efficiency of the electrochemical process could be justified. Simulation studies for the dielectrophoresis-adsorption process This work introduced the ap- plication of dielectrophoresis-forces to enhance the adsorption of PFAS. However, computational simulations are necessary to quantify the extent to which this enhancement is based on the dielec- trophoretic force effect. In addition, studies with multiple PFAS molecules should be conducted to determine the feasibility of the process for other PFAS. Optimization of the material to use during the dielectrophoretic enhanced-adsorption process The rapid saturation of the carbon-coated electrodes during the dielectrophoretsis-enhanced ad- sorption process suggest the necessity to optimize the electrodes fabrication process and/or replace the adsorption material of use. 173