Plasma-Based Treatment of Water Effluents Contaminated with Polyfluoroalkyl Substances (PFAS): Degradation and Identification of the Products of the Destruction Process

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Plasma-Based Treatment of Water Effluents Contaminated with Polyfluoroalkyl Substances (PFAS): Degradation and Identification of the Products of the Destruction Process
Song, Meiting
University of Florida
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Master's ( M.S.)
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University of Florida
Degree Disciplines:
Environmental Engineering Sciences
Committee Chair:
Bonzongo,Jean-Claude J
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Committee Members:
Winner,Lawrence Herman
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PFASs (Per- and Poly-fluoroalkyl Substances) are emerging pollutants of concern, which have been detected in different environmental matrices on a worldwide scale. Because of their potential negative effects on human health, which include cancer, ongoing research aims at preventing human exposure by developing different water, air and soil remedial strategies. In this study, a plasma-based water treatment process is evaluated for the removal of PFOA (perfluorooctanoic acid) and PFOS (perfluorooctanesulfonic acid) as example PFASs, from contaminated waters. Laboratory experiments were conducted using water samples with different chemistries, spiked with PFOA or PFOS. A subset of used water samples was treated by plasma (treatment group) and analyzed comparatively with the non-treated group (or control group). Plasma-based degradation of PFOA and PFOS was investigated under two different water recirculating flow rates of 0.8 and 1.4 gallons per minute (GPM). Concentrations of the PFASs in tested water samples were determined by electrospray ionization double mass spectrometry (ESI-MS/MS) in either negative or positive mode. The obtained results confirmed the high efficiency of PFASs degradation by plasma-based techniques. Briefly, the determined optimum conditions for plasma use (1.4 GPM) resulted in 95.25% and 100% degradation of initial PFOS and PFOA concentrations, respectively. Next, a study on the determination of potential breakdown products was initiated. Solutions of pure water spiked with PFOA and PFOS were treated by plasma to help identify the potential breakdown products in a simple water matrix. Preliminary results indicated the presence of degradation products, characterized by either the presence of new ion-peaks in the mass spectra or the increase of the surface areas of ion-peaks present in controls. Analyses related to the determination of the chemical structures of these degradation compounds is ongoing. Finally, when compared to results available in peer-reviewed literature, the plasma-based water treatment process tested in this study performs better microbial, sonochemical, and advanced oxidation-based techniques.

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© 2019 M eiting S ong


To my family


4 ACKNOWLEDGEMENTS I would like to thank my advisor, Dr. Jean Claude Bonzongo for his guidance and trust during my journey as graduate student at the University of Florida . His advices and counseling inspired me and allowed me to give my best in both laboratory experiments and the writing of this thesis . I thank both Dr s. Michael Annable and Subrata Roy for serving as Committee Members feedback and suggestions to help improve my work . I particularly thank Dr. Roy and Mr. Alexander Schindler Tyka for access to their plasma based treatment equipment and setup . Many thanks to the generous staff I met at the water reclamation facilities located on UF campus, Kanapaha , and Main Street in Gainesville, Florida , whom provided with waste water effluents used in my experiments . I would like to thank Dr. Townsend for provid ing us with landfill leachate samples used in this study . With regard to the PFAS analysis, I thank Dr. Jodie Johnson from the Chemistry Dept. for the different analyses using LC MS/MS technologies. I also thank Chenyi Yang who worked side by side with me in most of my lab experiments, as well as Zi chen Zhou, Lang Zhou and Qasem Alhadidi for being good officemates and friends.


5 Finally, I would like to thank my parents and my sister. They g a ve me both encouragement and spiritual support and ma d e me feel confident about myself. Their support has been crucial in helping me complete this thesis .


6 TABLE OF CONTENTS page ACKNOWLEDGEMENTS ................................ ................................ ................................ .. 4 LIST OF TABLES ................................ ................................ ................................ ............... 8 LIST OF FIGURES ................................ ................................ ................................ ............. 9 LIST OF ABBREVIATIONS ................................ ................................ .............................. 11 ABSTRACT ................................ ................................ ................................ ....................... 14 CHAPTER 1 POLYFLUOROALKYL SUBSTANCES IN THE ENVIRONMENT ............................ 16 1.1 Background and Problem Statement ................................ ................................ ... 16 1.2 Classification, Structure and Application ................................ .............................. 19 1.3 Bioaccumulation, Toxicity and Health Effects ................................ ...................... 21 1.4 Regulations and Rules ................................ ................................ ......................... 24 1.5 Occurrence and Fate of PFASs in Different Environmental Compartments ....... 26 1.5.1 PFASs in Raw and Drinking Water ................................ ........................... 26 1.5.2 PFASs in Wastewater and Sewage Sludge ................................ .............. 27 1.5.3 PFASs in Landfills ................................ ................................ ...................... 30 1.6 Existent Methods to Remove PFASs ................................ ................................ ... 33 1.6.1 Traditional Techniques ................................ ................................ .............. 33 1.6.2 Coagulation ................................ ................................ ................................ 33 1.6.3 Granular Activated Carbon ................................ ................................ ........ 34 1.6.4 Advanced Oxidation Processes ................................ ................................ . 35 1.6.5 Anion Exchange ................................ ................................ ......................... 36 1.6.6 Reverse Osmosis ................................ ................................ ....................... 37 1.6.7 Sonochemical Destruction ................................ ................................ ......... 38 1.6.8 Photolysis ................................ ................................ ................................ ... 40 1.7 Research Hypothesis ................................ ................................ ............................ 41 2 PLASMA BASED WATER TREATMENT FOR THE DEGRADATION OF PERFLUOROALKYL SUBSTANCES IN CONTAMINATED WATERS ................... 52 2.1 Introduction ................................ ................................ ................................ ........... 52 2.2 Material and Methods ................................ ................................ ........................... 56


7 2.2.1 Water Samples Collection ................................ ................................ ......... 56 2.2.2 Reagents, Standards and Characterization of the Tested Water Samples ................................ ................................ ................................ ........... 57 2.2.3 Water Sample Spiking with PFOA and PFOS ................................ .......... 57 2.2.4 Plasma Water Treatment ................................ ................................ ........... 58 2.2.5 Extraction and Analysis of PFOA and PFOS ................................ ............ 59 2.3 Results and Discussions ................................ ................................ ...................... 61 2.3.1 Chemical Characterization of the Different Water Samples Used in this Study ................................ ................................ ................................ .......... 61 2.3.2 Detection of PFOA and PFOS ................................ ................................ ... 62 2.3.3 Effect of Plasma based Treatment Procedure on the Degradation of PFOA and PFOS ................................ ................................ ............................. 63 2.3.4 Optimization of the Degradation of PFOA and PFOS .............................. 63 2.3.5 Identification of the Degradation Products of PFOA and PFOS Following Plasma Treatment. ................................ ................................ .......... 65 3 SUMMARY AND CONCLUSIONS ................................ ................................ ............ 77 3.1 Summary of Research ................................ ................................ ......................... 77 3.2 Future Work Recommendations ................................ ................................ ......... 77 REFERENCES LIST ................................ ................................ ................................ ........ 79 BIOGRAPHICAL SKETCH ................................ ................................ ............................... 96


8 LIST OF TABLES Table page 1 1 PFAS Reported in Recent Literature Together with Their Acronyms and Structures. Adapted from Jahnke and Berger (2009). ................................ ......... 43 1 2 Summary of PFASs Concentrations Detected in Influent and Secondary Treated Wastewater (as ng/L), and in Sewage Sludge (as ng/g) in Sewage Treatment Plants worldwide. Adapted from Arvaniti and Stasinakis (2015). ...... 45 1 3 Concentration (ng/L ± 95%CI) of Fluorochemical Analytes in Leachate from Six Landfill Leachates (A D) and a Laboratory Bioreactor. Adapted from Huset et al. (2011). ................................ ................................ ................................ 47 2 1 Chemical characterization of the different water samples tested. Samples labeled WWTP were collected from wastewater treatments plants in 3 different locations in the city of Gainesville, Florida. ................................ ............. 68


9 LIST OF FIGURES Figure page 1 1 Timeline of the production, commercialization and legislation of perfluoroalkyl carboxylic acids (PFCAs; at the top) and perfluoroalkyl sulfonic acids (PFSAs; at the bottom). Adapted from Hamid et al. (2018). ................................ 48 1 2 Typical average concentrations of perfluorooctane sulfonic acid and perfluorooctanoic acid in the blood (serum/plasma) from various countries. Adapted from Je nsen and Leffers (2008). ................................ ............................ 49 1 3 Persistent organic pollutants in blood plasma from pregnant women living in the Norwegian and Russian Arctic. Adapted from Jensen and Leffers (2008). .. 50 1 4 A brief history of PFAS production and regulation. Adapted from Kraft and Riess (2015). ................................ ................................ ................................ ......... 51 2 1 The process of ROS Water Treatment System. ................................ ................... 69 2 2 Illustration of example spectra determined using sample WWTP 2 and analysis by the negative electrospray ionization mass spectrometry (abbreviated ( )ESI MS). ................................ ................................ ....................... 70 2 3 Effects of two water treatment procedures on the removal efficiency of PFOA and PFOS. ................................ ................................ ................................ ............. 71 2 4 Chromatograms of the degradation PFOA using WWTP 1 water as an example, with the plasma non treated (top) and treated (bottom) shown comparatively. ................................ ................................ ................................ ........ 72 2 5 Chromatograms of the degradation PFOS using WWTP 1 water as an example, with the plasma non treated (top) and treated (bottom) HPLC/( )ESI MS mass chromatograms of the sum of the m/z 499 [M H] and the m/z 1021 [2M 2H+Na] . ................................ ................................ ................................ 73 2 6 Comparative performances of different PFASs degradation techniques based on removal efficiencies published in the literature and data from this study. ...... 74 2 7 Potential degradation products with m/z 272 ion peaks present in the plasma treated DI water containing either PFOA (c) or PFOS (d). ................................ .. 75


10 2 8 Chromatograms showing the (+)ESI MS ion peak with an m/z of 328.7 detected in plasma treated water (b) and in the plasma treated water samples containing PFOA and PFOS . ................................ ................................ . 76


11 LIST OF ABBREVIATIONS AFFFs Aqueous Film Forming Foams AOPs Advanced Oxidation Processes BDE Bond Dissociation Energy DAF Dissolved Air Flotation DBD Dielectric Barrier Discharge DI water Deionized Water DiPAPs Di substituted Fluorotelomer Phosphate Esters DiSAmPAP EtFOSE based Polyfluoroalkyl Phosphate Diester DOC Dissolved Organic Carbon ECF Electrochemical Fluorination ECHA European Chemicals Agency EFSA European Food Safety Authority EQS Environmental Quality Standard EtFOSAA Ethylperfluorooctane Sulfonamide Acetic Acid EtFOSE Ethyl perfluorooctane Sulfonamidoethanol FASAs Perfluoroalkyl Sulfonamides FTOHs Telomer Alcohols FAAs Fluoroacetamide GAC Granular Activated Carbon GPM Gallon Per Minute ISCO In Situ Chemical Oxidation LL Landfill Leachate


12 LOQ Low Detection Quantifications MSREC Mass Spectrometry Research and Education Center MSW Municipal Solid Waste MW Molecular Weight NCBI National Center for Biotechnology Information NF Nanofiltration nM Nanomolar NNRL New River Regional Landfill n:2 FTCA n:2 Fluorotelomer Carboxylic Acids n:2 FTUCAs n:2 Unsaturated Fluorotelomer Carboxylic Acids n:3 FTCAs n:3 Fluorotelomer Carboxylic Acids n:2 FTSAs Fluorotelomer Sulfonates PAC Powdered Activated Carbon PAPs Polyfluoroalkyl Phosphate Esters PBT Persistence, Bioaccumulation and Toxicity PerFASs Perfluoroalkyl Substances PFAAs Perfluoroalkyl Acids PFAS Per and Poly fluoroalkyl Substances PFBA Perfluorobutanoic Acid PFBS Perfluorobutanesulfonate PFCAs Perfluoroalkyl Carboxylates PFCs Perfluoroalkyl Compounds PFHpA Perfluoroheptanoic Acid


13 PFHxA Perfluorohexanoic Acid PFHxS Perfluorohexanesulfonate PFNA Perfluorononanoic acid PFOA Perfluorooctanoic Acid PFOS Perfluorooctanesulfonic Acid PFOSA Perfluorooctanesulfonamide PFPeA Perfluoropentanoic Acid PFSAs Perfluoroalkyl Sulfonates PolyFASs Polyfluoroalkyl Substances ppm Parts Per Million ppt Parts Per Trillion pTDI Provisional Tolerable Daily Intakes RO Reverse Osmosis ROS Reactive Oxygen Species SCID Source Collision induced Dissociation SNUR Significant New Use Rule STPs Sewage Treatment Plants SVHC Substances of Very High Concern USEPA United States Environmental Protection Agency vPvB Very Persistent, Very Bioaccumulative WWTPs Wastewater Treatment Plants


14 Abstract of Thesis Presented to the Graduate School of the University of Florida in Partial Fulfillment of the Requirements for the Degree of Master of Science PLASMA BASED TREATMENT OF WATER EFFLUENTS CONTAMINATED WITH POLYFLUOROALKYL SUBSTANCES (PFAS): DEGRADATION AND IDENTIFICATION OF THE PRODUCTS OF THE DESTRUCTION PROCESS By M eiting S ong M ay 2019 Chair: Jean Claude J. Bonzongo Major: Environmental Engineering Sciences PFASs ( Per and Poly fluoroalkyl Substances ) are emerging pollutants of concern, which have been detected in different environmental matrices on a worldwide scale. Because of their potential negative effects on human health, which include cancer, ongoing research aims at preventing human exposure by developing different water, air and soil remedial strategies. In this study, a plasma based water treatment process is evaluated for the removal of PFOA (perfluorooctanoic acid) and PFOS (perfluorooctanesulfonic acid) as example PFASs, from contaminated waters. Laboratory experiments were conducted using water samples with different chemistries, spiked with PFOA or PFOS. A subset of used water samples w as treated by plasma (treatment group) and analyzed comparatively with the non treated group (or control group). Plasma based degradation of PFOA and PFOS was investigate d under two different water recirculating flow rates of 0.8 and 1.4 gallons p e r minute (GPM).


15 Concentrations of the PFASs in tested water samples were determined by electrospray ionization double mass spectrometry (ESI MS/MS) in either negative or positive mode. The obtained results confirmed the high efficiency of PFASs degradation by plasma based techniques. Briefly, the determined optimum conditions for plasma use (1.4 GPM) resulted in 95.25% and 100% degradation of initial PFOS and PFOA concentrations, respectively. Next, a study on the determination of potential breakdown products was initiated. Solutions of pure water spiked with PFOA and PFOS were treated by plasma to help identify the potential breakdown products in a simple water matrix. Preliminary results indicated the presence of degradation products, characterized by either the presence of new ion peaks in the mass spectra or the increase of the surface areas of ion peaks present in controls. A nalyses related to the determination of the chemical structures of these degradation compounds is ongoing. Finally, when compared to results available in peer reviewed literature, the plasma based water treatment process tested in this study performs better microbial, sonochemical, and advanced oxidation ba sed techniques.


16 CHAPTER 1 POLYFLUOROALKYL SUBSTANCES IN THE ENVIRONMENT 1.1 Background and Problem Statement Per and poly fluoroalkyl substances (PFAS) are man made chemicals categorized as emerging contaminants, but which are now detected in the environment at concentrations ranging from subparts per trillion (ppt) to parts per million (ppm) levels (Thompson et al . 2011, Zareitalabad et al. 2013) . The y are characterized by a hydrophobic alkyl chain saturated by fluorine atoms, 4 to 18 carbons, and a hydrophilic functional group ( Prevedouros, Cousins and Buck 2006 ) . The magnificent electronegativity of fluorine creates an electrostatic attraction between C and F atoms, forming strong bonding property (Hagan, Chambers and He 2008) . Specifically, the bond dissociation energy (BDE) of a polarized carbon fluorine bond is approximately 544 kJ/mol, and believed to be one of the stronges t bonds reported in chemistry (Lemal 2004) . Due to their peculiar physicochemical characteristics, which include high chemical and thermal stability, and high water solubility ; PFASs contain ing products have been widely used in commercial and industrial applications since the 1950s (Buck et al. 2011, Hamid, Li and Grace 2018, Jahnke and Berger 2009, Thompson et al. 2011) . Other than the manufacturing of more resistant and durable products, the aqueous surface tension reduction property of PFASs have also been used in coatings, aqueous film forming foams (AFFFs) production, and household products such as non -


17 stick cookware and waterproof carpets (Jahnke and Berger 2009, Zareitalabad et al. 2013) . However, nearly 80% of perfluoroalkyl carboxylates (PFCAs) released to the environment can be related to direct contamination during fluoropolym er manufacture and use (Prevedouros, Cousins and Buck 2006, Wang and Chen 2009) . Numerous subsets of PFASs chemicals with various chain lengths and branched chemical groups have been detected in different environmental matrices. Among them, perfluoroalkyl acids (PFAAs) consisted of perfluoroalkyl sulfonate (PFSA, F(CF 2 ) n SO 3 ) and perfl uoroalkyl carboxylate (PFCA, F(CF2) n CO 2 ) are most studied (Rahman, Peldszus and Anderson 2013) . PFAA precursors like perfluoroalkyl sulfona mides (FASAs) and telomer alcohols (FTOHs) drew increasing attention since their degradation compounds can form additional PFAAs (Stock and Furdui 2007, Xia et al. 2006) . Also, ow ing to the complexity and occurrence of several isomers, it has been rather challenging to fully understand the fate and behavior of individual PFAS in the environment (Rahman et al. 2013) , except for the frequently targeted PFAAs like PFOA and PFOS. The extreme resistance of PFAS to natural degradation and to both thermal or biological breakdown processes (Liou et al. 2010, Webster 2008, Goss 2007) , have allowed long distances dispersal through air and water, making them ubiquitous on Earth.


18 The determination of the potential of PFAS to bioaccumulate and induce toxicity raised concerns for both the environment and human health. In fact, PFASs have been detected not only in drinking water at ppt levels ( Boulanger et al. 2005, Sasaki et al. 2002) , but also in human tissues (Bartell et al. 2010) . Risk assessment studies show that exposure to PFAS could lead to adverse health effects such as cancer, elevated cholesterol, obesity, suppression of immune system, and endocrine disruption (Barry, Winquist and Steenland 2013, Braun et al. 2016, Grandjean and Andersen 2012) . Concerns on human exposure are mostly for PFOA and PFOS, and would occur primarily through drinking of contaminated water and ingestion of tinted crops (Brooke et al. 2004) . Accordingly, PFASs have been regarded as global pollutants (Labadie et al. 2011) , and most studies on monitoring, fate and biological impacts of PFASs have been on two eight carbon members: PFOA and PFOS. In the early 2000s, the production of PFOS and analogs was phased out in the U.S. and Europe, while substitutes like PFAAs with shorter chain length were still manufactured elsewhere (Thompson et al. 2011) . Currently, a maximum concentrat ion of 70 ng/L for individual or combined PFOA and PFOS have been recommended by the U S EPA (the United States Environmental Protection Agency) as a lifetime health advisory for drinking water (USEPA 2016) . Regulation levels of PFOA and PFOS of 70 ng/L in drinking water have also been employed by several European countries


19 (McCleaf et al. 2017) . In Germany, a health based precautionary value of 0.1 /L is adopted for both compounds (Commission et al. 2006) . 1.2 Classification, S tructure and A pplication PFASs are defined as aliphatic substances, which contain one or more carbon atoms where H atoms (except the H atoms attached on the carbon atoms in functional groups) are substituted by fluorine (Buck et al. 2011) . In general, the chemical formulas of PFASs contain the fragment C n F 2n+1 (Buck et al. 2011) , in addition to the functional groups, which are characteristics to each compound. Therefore, PFAS consist of a hydrophobic alkyl chain and usually possess a hydrophilic fu nctional group. They are commonly divided into 3 classes: the perfluoroalkyl substances (PerFASs), polyfluoroalkyl substances (PolyFASs), and fluorinated polymers (Jahnke and Berger 2009, Codling et al. 2014) . The per and poly fluorinated alkyl substances (PFAS) reported in recent literature (Table 1 1) have physicochemical properties that depend on the length of the chain and type of functional group (Wang et al. 2011, Ding and Peijnenburg 2013) . Overall, per fluoro chemicals (organics with hydrogens replaced by fluorines) are extremely recalcitrant and environmentally persistent (Vecitis et al. 2009) . Fluorinated polymers include a large variety of substances that can also be sub divided into three groups: (i) the fluoropolymers, (ii) the per fluoropolyethers, and (iii) the side chain fluorinated polymers (Codling et al. 2014, Buck et al. 2011) . The fluorochemicals are hard to treat or remove by most conventional technologies (Vecitis


20 et al. 2009, Sinclair 2006, Higgins et al. 2006) . The perfluorinated carbon tail of perfluorinated surfactants prefer to separate from the aqueous phase, and an ionic headgroup tends to separate into the aqueous phase. Due to their biphasic or surfactant nature, they preferentially accumulate at the interface of air and water ( Shinoda et al. 197 2 , Lo 2005, Simister et al. 1992, Lu, Ottewill and Rennie 2001) . The extreme ly stable and strong carbon fluorine bond (Smart and Tatlow 1994) , as well as the highly polar property, contribute to the peculiar chemical and thermal stabilities of some perfluoroalkyl compounds and to their hydrophobic and lipophilic nature (Kissa 2001, Kissa 1994, McCleaf et al. 2017) . Two application processes of PFAS production, electrochemical fluorination (ECF) and telomerization have been described (Jahnke and Berger 2009) . Since 1947, the ECF process has been an application to produce isomer mixtures mainly for the linear isomer including a number of branched isomers. Nevertheless, by the end of 2002, a company called 3M phas ed out its C 8 based production although 3M was the only major company known to use ECF process to produce PFASs (Renner 2001) . In which case, due to the similar properties, the C 8 product line has been partly replaced by C 4 analogs (Jahnke and Berger 2009) . Telomerisation is another main process to produce PFASs. Since the 1970s, it has been commercially used and exclusively produced almost linear isomers, such as perfluoroalkyl 2 ethanols or the partly fluorinated fluorotelo mer alcohols (FTOHs) (Kissa 2001) .


21 used in numerous industria l and commercial applications (Kissa 2001) . The physicochemical properties of PFAS have been used mostly in the surface treatment of textiles, c arpet and leather, food packaging, paper, performance chemicals (such as AFFF s , and herbicides/insecticides or emulsifiers in fluoropolymer production ). Figure 1 1 shows a timeline of the production, commercialization and legislation of perfluoroalkyl carboxylic acids and perfluoroalkyl sulfonic acids (Kissa 2001, Prevedouros et al. 2006, Hamid et al. 2018) . 1.3 Bioaccumulation, T oxicity and H ealth E ffects Perfluorinated substances are classified as emerging contaminants due to their global distribution, persistence and bioaccumulation potential hazards. Numerous ionic PFASs resist metabolism, photolysis, hydrolysis and microbial degradation (Kissa 2001) . The persistence, bioaccumulation and distribution of PFASs throughout the food health and the environment (Giesy and Kannan 2001, Lau et al. 2007, Martin et al. 2003) . In contrast to ionic species, neutral PFAS s are usually not persistent in the environment. Nevertheless, they are carri ed out long distances through the atmosphere, thanks to their high volatility. However, these polyfluorinated compounds have been considered as precursors of the persistent ionic PFAS (Hekster et al. 2002, Tomy et al.


22 2004, Ellis et al. 2003, Ellis et al. 2004, Martin, Mabury and O'Brien 2005, Martin et al. 2006, D'Eon et al. 2006) . The ecotoxicity of these compounds has also been investigated, but details on ecoto xicology are still needed. Wang et al. have reported that the mixtures of PFASs have a toxic effect on Photobacterium phosphoreum (Wang et al. 2011) . On the other reported signs of increased ecotoxicity in target binary mixtures as the molar ratio of PFOS increased (Liu et al. 2010, Liu et al. 2009, Liu, Du and Zhou 2007) . These studies suggested that the mixtures of PFAS can result in unpredictable effects. Howev er, the potential impacts of PFASs mixtures need to be systematically assessed to determine whether and how these chemicals collectively affect the integrity of aquatic ecosystems. Perfluoroalkyl compounds (PFCs) may through wastewater effluent, rain, sep tic discharge and runoff enter the environment or via the application of sludge to agriculture lands. Discharges of wastewater from municipal wastewater treatment plants (WWTPs) are recognized as one of the significant point sources of PFCs to aquatic envi ronments (Boulanger et al. 2005, Schultz, Barofsky and Field 2006a, Sinclair 2006) . As a consequence of the widespread use of PFASs and their resulting emissions, a broad range of these substances has been detected in the environment, wildlife, and hum ans. The global extent of such contamination was first demonstrated for perfluorooctane sulfonic acid, C 8 F 17 SO 3 H (PFOS) in wildlife (Giesy and Kannan


23 2001) . At about the same time, it was reported that PFOS, PFOA, and other PFASs were present in numerous samples of human blood (Giesy and Kannan 2001) . The first reports on high PFOS le vels in blood were from workers in perfluorochemical manufacturing (Olsen et al. 1999) , and from contaminated groundwater from a fire training site (Moody and Field 1999) . A publication on the ubi quity of PFOS in wildlife around the world followed shortly afterward, raising concern among scientists, regulators and industry alike (Giesy and Kannan 2001) . In humans, PFAS are readily absorbed and bind to proteins in blood serum (Figure 1 2). They accumulate mainly in organs such as liver, kidney and spleen, but also in testicles and brain (Heuvel, Kuslikis and Peterson 1992, Austin et al. 2003, Jones et al. 2003) . While the acute toxicity of PFAS is moderate, it increases with the carbon chain length. The oral rat LD50 for PFOS is 251 mg / kg (USEPA, 2000), while the oral LD50 for PFOA is between 430 and 1800 mg / kg (Kennedy et al. 2004) . In fact, exposure to polyfluorinated substances measured on basis of whole blood may be the known environmental contaminants, brominated flame retardants and even phthalates (Figure 1 3). The results of a bioaccumulation study conduct in Sweden showed that high local PFASs exposures should be of concern (Borg et al. 2013) . Also, a study was conducted to try to link PFAS and PFOA to cancer using a population of 2,500 males


24 and females suffering from 21 d ifferent types of cancer (Barry et al. 2013) . In the study, a cohort comprised of adults who worked at a local chemical plant or lived around contaminated water districts in Mid Ohio Valley was selected. They measured the serum PFOA of most participants during a survey conducted in 2005 2006, and then they obtained further medical history in 2008 2011. After recording yearly PFOA serum concentrations for each participant from 1952 through 2011 and validation and review self reported cancers were validated through medical records and cancer registry respectively, the estimation of connection between cancer and cumulative serum concentration of PFOA by proportional hazards models. The conclusion of analysis from Barry indicated that PFOA exposure had a positive relationship with kidney and testicular cancer from this cohort (Barry et al. 2013) . 1.4 Regulations and R ules Based on the adverse effects of PFAS mentioned of PFAS , a set of rules and regulations have been proposed and summarized in Fig ure 1 4 . As mentioned previously, since 2000, the 3M company has phased out the production of PFOA, PFOS and related chemicals under the guidance of the United States Environmental Protection Agency (USEPA). Also, the USEPA enacted a Stewardship Program including 8 major chemical companies in the U.S., Japan and Western Europe to reduce by 95% the emissions of PFOA precursors from facilities that degrade to PFOA and related analogs by 2010; and to commit to eliminate emissions and production of these compounds by 20 15 (USEPA


25 2015a) . For PFOS, the World Semiconductor Council guided the virtual elimination of PFOS using semi conductor production (World Semiconductor 2016) . In 2008, based on the animal studies, the European Food Safety Authority (EFSA) proposed provisional tolerable daily intakes (pTDI) values of 1500 ng/kgbw per day for PFOA and 150 n g/kgbw per day for PFOS (Alexander et al. 2008) . Since 2009, POSF, PFOS and their related chemicals such as their salts, higher homologues are listed as very persistent, very bioaccumulative (vPvB) and toxic compounds in Annex B in the Stockholm Convention on Persistent Organic Pollutants, indicating they have been strictly restricted on their production and use (Krafft and Riess 2015, UNEP 2009, Persistent Organic Pollutants Review) . However, production and use for acceptable applications were recognized by the conventi on and other regulations. On the other hand, it should be mentioned that there were unacceptable chemicals, such as PHxSF derivative s , because of their potential to act as precursors of long half life compounds. In 2012, the very long F chain compounds of PFCAs, C 11 C 14 , have been classified as potentially hazardous contaminants of vPvB and recorded in the Candidate List of Substances of Very High Concern (SVHC) (ECHA 2014) . PFOA and the relevant ammonium salts obviously meet the standards of persistence, bioaccumulation and toxicity (PBT) substances (Vierke et al. 2012) . They were identified by the European Union, and as the reproductive toxicity, PFOA and APFO joined the SVHC list in 2013


26 (ECHA 2014) . Additionally, PFOA was classified as a potential carcinogenic to human by the working group of the International Agency for Research on Cancer in 2014 (Benbrahim T allaa et al. 2014) . In the United States, F polymers are no longer exempted from the notifications of pre production or importation (Matthews and Solomon 2005) . In 2015, a proposed rule in the United States, Significant New Use Rule (SNUR), states that the phased PFASs wou ld not re enter the market without review (USEPA 2015b) . Following a study of risk assessment, PFOS was included in the list of priority haza rdous substances by the European Commission recently. Therefore, it should be monitored in European Union water bodies, at the level of the Environmental Quality Standard (EQ S ) of 0.65 ng/L f or freshwaters. Whereas, USEPA and NCBI ( National Center for Biot echnology Information ) recommended, in drinking water, the Provisional Health Advisories of 400 ng/L for PFOA and 200 ng/L for PFOS (USEPA 2 009) . Regulatory restrictions on the use of PFOA and PFOS have led major PFAS manufacturers to seek alternatives to these compounds, especially homologues with different chain lengths. 1.5 Occurrence and Fate of PFASs in Different Environmental Compartments 1.5.1 PFASs in Raw and Drinking W ater Compared to surface and ground waters, there were fewer reports on the occurrence and fate of PFASs in both raw and drinking waters. When investigated,


27 studies on PFASs in raw and drinking wa ters focused primarily on PFOS and PFOA, and these two chemicals are the most reported. Nevertheless, some recent studies from Europe have reported the concentrations of shorter chain PFASs in drinking water with even higher concentrations than PFOA and/or PFOS at some locations. Examples include PFBA (Perfluorobutanoic Acid) , PFBS and PFHxA (Ullah, Alsberg and Berger 2011, Esch auzier et al. 2012) . PFBS and PFBA were detected at the average concentrations of 20 ng/L and 30 ng/L, respectively, showing the highest monitored PFASs in samples which were collected from a treatment plant in Amsterdam (Eschauzier et al. 2012) . Branched isomers of PFOA and PFOS were also detected in drinking water samples (Eschauzier et al. 2012) . 1.5.2 PFASs in Wastewater and Sewage Sludge In the past 10 years, numerous monitoring data for PFCs occurrence and fate in influent s and tr eated wastewater and sew age sludge have been published (Boulanger et al. 2005, Sinclair 2006, Schultz et al. 2006a, Guo et al. 2010, Becker, Gerstmann and Frank 2008, Huset et al. 2008, Ahrens 2011) . In a review paper published in 2015, 22 different PFCs (C 4 C 14 , C 16 , C 18 carboxylates, C 4 C 8 and C 10 sulfonates and 3 sulfonamides) were observed in urban and/or industrial wastewater samples (Arvaniti and Stasinakis 2015) . Most of the data from the monitoring studies were from the US, Northern Europe and Asia , but limited information is available for both Canada and Australia. In untreated wastewater and


28 sludge, the concentration of PFC S can reach hundreds of ng/L and thousands of ng/g dry w eight, respectively. In secondary biological treatment, they are not significantly removed, and their concentrations in treated wastewaters are usually higher than in untreated wastewaters. Their biodegradation in wastewater treatment seems impossible; how ever, recent studies have shown that precursor compounds may be converted to PFC during wastewater treatment. The adsorption of sludge by PFC S has been deeply studied, and it seems to be an important mechanism to control the removal of STP s (Sewage Treatme nt Plants) sludge. In tertiary treatment technology, significant PFC s removal has been observed using activated carbon, nanofiltration, reverse osmosis or advanced oxidation and reduction processes. Most studies are conducted in p ure water, and in many cases, experiments are carried out under extreme laboratory conditions ( high concentration, a high radiation sou rce, temperature or pressure). Therefore, there is a need for work focus ing on the understanding of bioconversion processes in aerobic and anaerobic bioreactors, PFC s formation, and the application of advanced treatm ent technologies u nder STP conditions . Rui Guo's study has shown that one significant source of PFCs in natural water is WWTP s (Guo et al. 2010) . In this study , 10 different PFCs chemicals were measured in samples from wastewater influent, effluent and sludge from 15 different municipal ities, 4 live stock and 3 industrial WWTPs. Table 1 3 shows the concentrations of measured PFCs. The c oncentration of PFOS in sludge samples ranged from 3.3 to 54.1 ng/g.


29 PFOA was the major compound in wastewater , with concentrations ranging from 2.3 to 615 ng/L and 3.4 to 591 ng/L in influent s and effluent s, respectively . The fate of PFCs in wastewater treatment plants were found to be related to the presence of specific functional groups. After treatment , in most WWTPs, PFOA concentrations appeared to increase whereas that of PFOS decreased. The normalized partition coefficient of organic carbon of PFASs was higher than that of carboxylate analogues, suggesting a partition tendency of PFASs between water and sludge. The high concentration of PFCs was contained by industrial WWTPs, however, due to their low emissions, these WWTPs we re not the main o rigin of PFCs chemicals in the studied water system . There were more PFCs discharged from sewage treatment plants in big cities, which indicated that domestic sewage was one of the important sources of PFC pollution. The occurrence and fat e of 9 different perfluoroalkyl carboxylic acids and 3 perfluoroalkyl sulfonic acids in Italy's most industrialized areas were investigated (Castiglioni et al. 2015) . S amples were collected from WWTPs, major ri vers flowing through the basin, and from untreated groundwater and finished drinking water. In samples from WWTPs receiving industrial wastes, perfluorinated substances were not removed, and the disc harge volume was up to 50 times compared to a load of wastewater treatme nt plants receiving urban wastewater .


30 1.5.3 PFASs in Landfills At the end of the service lives, many various consumer products containing PFAAs and their precursors such as paper, textiles, carpets, packaging and other goods are sent to municipal landfills for disposal . Biological solids which contain chemicals of PFASs are also landfilled in many municipalities (Guerra et al., 2014; Arvaniti et al., 2012). PFASs from the was te through the way of biological and abiotic leaching such as desorption are released (Allred et al., 2015). PFCAs with 4 14 carbon chain length and PFSA of mostly even chain length from C 4 C 10 have been reported in landfill leachate in the ng/L to mg/L r ange (Hamid et al. 2018) . Possible sources of PFAAs include consumer products (e.g., paper, textile, packaging, food contact paper, carpet) Zushi and Masunaga 2015) , building materials , electronics resulting from intentional addition of PFAAs during production and/or product use, and contamination with by pr oducts impurities during production . Furthermore, PFAA precurs ors (e.g., FTOH, n:2 fluorotelomer carboxylic acids (n:2 FTCA) and n:2 unsaturated fluorotelomer carboxylic acids (n:2 FTUCAs)) present in the consumer products (Kotthoff et al. 2015, Ye et al. 2015) can degrade to PFAAs during product use and/or after disposa l in the landfill (Allred et al. 2015, Lang et al. 2016) . Meanwhile, Fluorotelomer based compounds such as, n:2 FTCAs, n:2 FTUCAs, n: 3 fluorotelomer carboxylic acids (n:3 FTCAs), fluorotelomer


31 sulfonates (n:2 FTSAs) have been detected in landfill leachate (Lang et al. 2017, Allred et al. 2015, Allred et al. 2014, Benskin et al. 2012, Huset et al. 2011) and lab scale landfill reactors (Lang et al. 2016, Allred et al. 2015) ranging from a few ng/L to mg/L. Several unsubstituted, methyl and ethyl perfluoroalkane sulfonamide acetic acids (FASAAs) with C 4 C 8 carbon chain length have been reported in landfill leachates (Lang et al. 2017, Allred et al. 2014, Benskin et al. 2012, Huset et al. 2011) . Biodegradation of ethyl perfluorooctane sulfonamidoethanol (EtFOSE), a major raw material of paper and packaging pr oducts (Buck et al., 2011), is said to form C8 based ethylperfluorooctane sulfonamide acetic acid (EtFOSAA) (Rhoads et al., 2008). Similar biodegradation pathways could be responsible for the shorter FASAA, MeFASAA and EtFASAA homologues resulting from met hyl and ethyl perfluoroalkyl sulfonamidoethanols (FOSE) (Allred et al. 2014) . In addition, detection of a few classes of polyfluoroalkyl phosphate esters (PAPs) (e.g., Di substituted fluorotelomer phosphate esters (6:2 10:2 DiPAPs ) and EtFOSE based polyfluoroalkyl phosphate diester (DiSAmPAP)) have been reported in leachate (Allred et al. 2014, Lang et al. 2017) . PAPs are used in papers and synthetic fibers to impart oil and water repellency, in semiconductor materials and in personal care products (Liu and Liu 2016) . Microbial degradation of PAPs resulting in a mixture of FTCAs and PFCAs has been reported in activated sludge Mabury 2010) , and in aerobic soil (Liu and Liu 2016, Lee et al. 2013) , accounting for the


32 infrequent detection of PAPs in leachate, despite their widespread use and high production volume (De Silva et al. 2012) . PFCAs are generally found to be the dominant PFASs (Fuertes et al. 2017, Allre d et al. 2014, Li et al. 2012, Huset et al. 2011) . Also, C 4 C 7 chain length PFAAs are more abundant than their longer chain (C 8 ) homologues (Fuertes et al. 2017, Li et al. 2012, Bossi et al. 2008, Kallenborn et al. 2004) . Short chain PFAAs are prone to preferential release and leaching from municipal solid waste (MSW), c onsistent with their higher aqueous solubilities and lower organic carbon water partition coefficients relative to longer chain PFAAs (Yan et al. 2015) . In the dominance of C 4 C 7 PFAAs could be related to the shift towards the production of shorter chain perfluorinated compounds since the early 2000s. An increasing number of studies showing degradation of polyfluorinated compounds to PFAAs in th e environment (Liu and Liu 2016) , along with increasing availability of chemical standards and improved analytical techniques, have led to recent studies (summarized in the next section) to investigate PFAA precursors and their degradation products, as well as other classes of perfluorinated compounds (e.g., perfluoroalkyl sulfonamide derivatives) in landfill leachate. Some of the fluorotelomer based (e.g., n:2 FTCA, FTUCAs) and N alkyl FASAAs are freque ntly detected with concentration ranges that are comparable to and/or higher than those of PFCAs (>mg/L).


33 1.6 Existent Methods to Remove PFASs 1.6.1 Traditional T echniques Conventional water treatment processes are not effective in removing PFASs. For inst ance, Zhang et al. (Zhang et al . 2011) found that PFASs are partly reduced by a drinking water treatment system in China where surface water was the source sample. Another study by Eschauzier et al. (Eschauzier et al. 2010) shows that riverbank filtration did not remove the PFASs. Skutlarek et al. (Skutlarek, Exner and Färber 2006) also clearly demonstrated that PFASs were not removed by water treatment steps and the concentrations in the surface waters co rresponded to PFAS levels in drinking water. Recently Quinones & Snyder (Quiñones and Snyder 2009) confirmed that the PFAS concentration of influent and effluent was similar in drinking water treatment plants in the USA, suggesting that treatment systems were ineffective in removing these c ompounds. Sand filtration and ozonation processes were also ineffective in removing PFOS and PFOA during drinking water treatment (Takagi et al. 2011) . A similar result was found by Thompson et al . (Thompson et al. 2011) in a water re clamation plant in So uth East Queensland, Australia. 1.6.2 Coagulation Deng et al. (Deng et al. 2011) found that coagulation (polyaluminium chloride) could remove the PFOA from water. This was because some PFOA transferred from the aqueous phase to a solid phase. They also found that the addition of powdered


34 activated carbon (PAC) before the coagulation process significantly enhanced PFOA rem oval efficiency. This was explained as the negative PFOS adsorbing onto the PAC via electrostatic interaction, resulting in the removal of PFOS and PFOA from water in the coagulation process. A similar observation was made by Yu et al. (Yu et al. 2009) that PAC could effectively remo ve the PFOA concentration through the electrostatic interaction and hydrophobic interaction between them. A greater coagula nt dosage (>60 mg/L) and lower pH (4.5 6.5) can enhance the PFAS removal (Xiao, Simcik and Gulliver 2013) . 1.6.3 Granular A ctiva ted C arbon The ability of wastewater treatment technologies to remove or degrade PFASs depends on the water treatment process. For instance, one study reported that granular activated carbon (GAC) could effectively remove PFOS from aqueous solutions (Ochoa Herrera and Sierra Alvarez 2008) . Another study did find that GAC removed PFOS and PFOA con centration s with an efficiency of 64 ± 11% and 45 ± 19%, respectively (Flores et al. 2013) . Takagi et al. reported that GAC could e ffectively remove the PFOS and PFOA when it was used for less than 1 year (Takagi et al. 2011) . In contrast, the effluent concentration of those compounds was higher than the influent concentratio n when it used for a longer time (>1 year), possibly due to growth of the biofilm into the pores and surface of the carbon, exhausting the adsorption capacity and possibly interaction between those compounds and between those compounds and the biofilm,


35 pro ducing lowered concentrations initially and higher concentrations after significant use. The adsorption capacity of the GAC is also related to the temperature of the water sample. Takagi et al. (Takagi et al. 2008) found that the removal of PFOA was 36 56% in summer and 31 58% in winter. The removal efficiency of PFOA and PFOS by GAC also depends on the volumes of activated carbon. Lampert et al. (Lampert, Frisch and Speitel Jr 2007) found in their batch test that more than 90% of both PFOA and PFOS were removed when the activated carbon was 0.1047 g or greater at 7 days of contact time. However, at a similar contact time, the reduction efficiency was decreased to a bout 50% for PFOA and 82% for PFOS when 0.0587 g GAC was used. Therefore, the use of GAC with greater volume and suitable regeneration regimes appear to be the important parameters in the efficient removal of PFASs. 1.6.4 Advanced O xidation P rocesses Adva nced oxidation processes (AOPs) that have been tested for PFAS removal include electrochemical oxidation, photolysis, and photocatalysis. During these processes, strong, oxidizing, and nonselective radicals are generated that can attack a variety of xenobi otics, such as pharmaceuticals (Ike hata, Jodeiri Naghashkar and Gamal El Din 2006) , phenols and dyes (Ahmed et al. 2011) , and trinitrotoluenes (TNTs) (Ayoub et al. 2010) . The focus of developing in situ chemical oxidation (ISCO) technologies for PFASs is AOPs. Common ISCO reagents include hydrogen peroxide, sodium


36 persulfate, potassium/sodium permanganate, ozone, and ozone/peroxide. The application of AOPs with oxidants suc h as hydrogen peroxide has yielded mixed results in laboratory settings. In addition, AOPs for groundwater treatment are often configured as ex situ technologies, reactions occur predominantly in the aqueous phase and, as a result, are more effective at tr eating dissolved phase contaminants than adsorbed compounds (H ao et al. 2014) . Chemical oxidation using AOPs has the potential to address PFAS source areas; however, to date, laboratory studies have provided mixed results, underscoring the importance of characterizing oxidant demand and the effect of AOP chemistry on the families of PFASs that often co occur with PFOA and PFOS. In addition, further research at the pilot scale is needed to demonstrate the ability to effectively implement these technologies in situ. 1.6.5 Anion E xchange Since PFOS exists as an anion in aqueous solution, anion exchange becomes a promising technique for its removal . B ut the previous work of Yu (Yu et al. 2009) demonstrated that the anion exchange resin (Amberlite IRA400) required over 168 h to achieve the sorption equilibrium for PFOS. Different resins vary signif icantly in some properties such as polymer matrix, porosity and a functional group, leading to the distinct removal efficiency for pollutants (Li and SenGupta 2000, Gu, Ku and Brown 2005, Tan and Kilduff 2007, Boyer, Singer and Aiken 2008) . Few studies were


37 conducted to compare the sorption of PFOS using different anion exchange resins (Lampert et al. 2007) , and the effect of resin characteristics on the sorption is unclear. 1.6.6 Reverse O smosis There are many studies demonstrating that the high pressure membranes such as those in nanofiltration (NF) and reverse osmosis (RO) can effectively remove the PFAS (Tang et al. 2006, Zhao et al. 2013) . Tang et al. (Tang et al. 2006) investigated the use of RO me mbrane in removing PFOS from wastewater with the concentration range of 0.5 to 1,500 mg/L and they found that >90% reduction was achieved by RO. Tang et al. (Tang et al. 2007) further studied the use of RO and NF to removal PFOS from wastewater and they got >99% and 90 99% removal for RO and NF membranes, respectively. Zhao et al. (Zhao et al. 2013) studied the NF membrane for removing PFOS from simulated surface water and found that the membrane can effectively reduce the concentration. They also investigated the effects of PFOS concentration, pH and calcium concentration on PFOS rejection. They found pH leads to an increase in the PFOS rejection when the pH increases from 3 to 9 which leads to an increase in the PFOS rejection from 86 to 95% and 93 to 97% in the presence of 0.1 mM Ca 2+ and 1 mM Ca 2+ at 0.4 MPa, respectively, because the increasing pH generally increases electrostatic interactions which play a role in P FOS rejection. Similarly, increasing calcium concentration decreased the permeate PFOS concentration because calcium ions bridge PFOS and the negatively charged membrane surface, which enhances the


38 adsorption of PFOS on the membrane. Another study shows th at average rejections of the PFASs were 99.3% for virgin RO membranes, but 95.3% for fouled RO membranes (Appleman et al. 2013) . The transmembrane pressure was not increased to maintain a et al. (Tang et al. 2006) recommended that high flux RO mem branes should be avoided when treating water with high concentrations of PFOS (>30 mg/L PFOS), these membranes normally have a low rejection effect and the advantage of a high flux cannot be maintained for a long time. 1.6.7 Sonochemical Destruction Sonochemical degradation of PFASs occurs through the application of ultrasound to an aqueous medium. When ultrasound is applied, cavitation bubbles form during the rarefaction (negative pressure) portion of sound waves (Thompson and Doraiswamy 1999, Joseph et al. 2009) . The cavitation bubbles will implode adiabatically, creating extreme temperatures (>9,700 in the vapor core) and pressures (14,000 psi) within its cavity (Didenko, McNamara and Suslick 1999, Ashokkumar and Grieser 2005, Ciawi et al. 2006, Eddingsaas and Suslick 2007, Park et al. 2009) . Highly reactive intermediates and radicals, including hydroxyl radicals, a hydrogen atom, and an oxygen atom, form during cavitation bubble collapse (Leighto n 2012) . This combination of highly reactive species and high temperatures and pressures has made sonolytic decomposition of PFASs successful. PFOS and PFOA


39 were completely mineralized through sonolysis to CO, CO 2 , F , and SO 4 2 , as detected by HPLC MS, ion chromatography, FT IR, and GC MS. Two studies report no detection of reaction intermediates and complete defluorination of PFOS within 3 h and PFOA within 2 h (Vecitis et al. 2008b) . There was immediate production of inorganic sulfur and fluorine atoms, with a slight delay in the production of CO and CO 2 (Vecitis et al. 2008b) . Complete mineralization was possible due to the presence of three different reactivity sites: inside the cavitation bubble, at the interfacial region between the cavitation bubble and the bulk aqueous solution, and in the bulk aqueous solution and vapor phase (Moriwaki et al. 2005, Vecitis et al. 2008b) . The sonolytic decomposition of PFASs depended on the type of gas used and the initial PFAS concentration. All sonolytic decomposition processes of PFASs have been conducted using argon gas since it will produce higher temperatures and increased reaction yields compared with air (Moriwaki et al. 2005) . However, Phan Thi et al. (Thi, Do and Lo 201 4) observed 100% PFOA decomposition when using nitrogen gas with NaHCO 3 . The initial concentration of PFASs is important as saturation kinetics could influence the reaction. At higher concentrations of PFOS or PFOA, zero order kinetics were observed, in dicating saturation of adsorption sites on the interfacial region, while at lower concentrations, pseudo first order kinetics took place (Vecitis et al. 2008a) . Sonolysis of PFASs may be improved in conjunction with other treatment methods, such as ozone, microwave irradiation, persulfate, and VUV (Yang et al. 2013 ) .


40 When sonolysis and ozone were applied to groundwater containing PFOS and PFOA, the degradation rates increased by 79% for PFOS and 70% for PFOA when compared with Milli Q water (Cheng et al. 2008) . Similarly, microwave irradiation combined with an ultrasonic homogenizer decomposed PFOA within only 90 s with 59% defluorination yields (Horikoshi et al. 2011) . Temperatures reac hed 1,000 at the tungsten tip and 51 in the bulk liquid. Active species were also generated, including hydroxyl radical, a hydrogen atom, and an oxygen atom, causing decarboxylation and oxidation of PFOA and its intermediates. Comparatively, a more ben ign treatment approach can be used with persulfate and sonolysis under either air or argon gas and has been used to degrade five perfluoroether carboxylic acids and two perfluoroether sulfonates (Hori et al. 2012) . Argon gas increased re moval yields compared with air due to the occurrence of higher temperatures when the cavitation bubbles collapsed. 1.6.8 Photolysis UV photolysis has been demonstrated to be effective at degrading PFOS and PFOA (Hori et al. 2004, Chen, Zhang and Liu 2007) . Fujii et al. (Fujii et al. 200 7) demonstrated that photocatalysis (reaction time up to 3 days) and advanced oxidation (with high temperature and pressure) could effectively degrade the PFOS and PFOA to CO 2 (Lampert et al. 2007) demonstrated more than 99% remo val of PFOS and PFOA using anion exchange (AIX) resins.


41 1.7 Research Hypothesis According to the statement above, PFASs are hazardous, bioaccumulative and toxic chemicals which have an adverse impact on human being. It is a priority to remove these compoun ds, even to degrade them. Moreover, as the typical and significant compounds, PFOA and PFOS have been detected, monitored and reported in numerous literature . However, there was not a method which can make them thoroughly degraded, but also is efficient, e conomic and easy operated . In this research, we focus on the degradation of PFOA and PFOS and identification their breakdown products in contaminated water after treatment by plasma based technique. According to the regulation of USEPA for PFOA and PFOS in drinking water for health and the LOD of liquid chromatography equipment, we decided to remove 100 ppb PFOA and 100 ppb PFOS from the collected water from different wastewater treatment effluents. Samples were classified into 2 groups, experimental group and control group, to compare the concentrations of PFOA and PFOS before and after the plasma based treatment. For the experimental group, we used 2 different methods to remove PFOA and PFOS in water samples. The main differences among these methods were the flow rates of pumping water. After the analysis of the relative concentrations in treated samples, we chose the experimental group which had the highest removal rate to measure the fragment s of PFOA and PFOS. In this step, due to the impurity of


42 wastew ater effluent, we used deionized water to replace wastewater effluent as the subject. Following the same concentrations (100 ppb) of PFOA and PFOS were spiked into deionized water , which would be treated by the group exhibited the highest removal rate, bre akdown products of PFOA and PFOS were analyzed by liquid chromatography again to ensure the exact substance. Before and after treatment, pH, alkalinities and ion concentrations of wastewater effluents and DI water were measured in order to analyze the environmental impact by plasma technology.


43 Table 1 1 . PFAS Reported in Recent Literature Together with Their Acronyms and Structures . Adapted from Jahnke and Berger (2009) . Compounds Acronym Structure Perfluoroalkane sulfonates (PFSAs) Perfluorobutane sulfonate PFBS CF 3 (CF 2 ) 3 SO 3 Perfluorohexane sulfonate PFHxS CF 3 (CF 2 ) 5 SO 3 Perfluorooctane sulfonate PFOS CF 3 (CF 2 ) 7 SO 3 Perfluorodecane sulfonate PFDS CF 3 (CF 2 ) 9 SO 3 Other sulfonates and sulfinates x:2 Fluorotelomer sulfonate x:2 FTS C x F 2x+1 CH 2 CH 2 SO 3 Perfluorohexane sulfinate PFHxSi CF 3 (CF 2 ) 5 SO 2 Perfluorooctane sulfinate PFOSi CF 3 (CF 2 ) 7 SO 2 Perfluoroalkyl carboxylates (PFCAs) Perfluorobutanoate PFBA CF 3 (CF 2 ) 2 COO Perfluoropentanoate PFPeA CF 3 (CF 2 ) 3 COO Perfluorohexanoate PFHxA CF 3 (CF 2 ) 4 COO Perfluoroheptanoate PFHpA CF 3 (CF 2 ) 5 COO Perfluorooctanoate PFOA CF 3 (CF 2 ) 6 COO Perfluorononanoate PFNA CF 3 (CF 2 ) 7 COO Perfluorodecanoate PFDA CF 3 (CF 2 ) 8 COO Perfluoroundecanoate PFUnA CF 3 (CF 2 ) 9 COO Perfluorododecanoate PFDoA CF 3 (CF 2 ) 10 COO Perfluorotridecanoate PFTrA CF 3 (CF 2 ) 11 COO Perfluorotetradecanoate PFTA CF 3 (CF 2 ) 12 COO Perfluoropentadecanoate PFPDA CF 3 (CF 2 ) 13 COO Fluorotelomer carboxylates x:2 Fluorotelomer carboxylate x:2 FTCA C x F 2x+1 CH 2 COO x:2 Fluorotelomer unsaturated carboxylate x:2 FTUCA C x 1 F 2x 1 CF=CHCOO Neutral PFAS x:2 Fluorotelomer olefin x:2 FTolefin C x F 2x+1 CH=CH 2


44 Table 1 1. Continued Compounds Acronym Structure x:2 Fluorotelomer alcohol x:2 FTOH C x F 2x+1 CH 2 CH 2 OH x:2 Fluorotelomer aldehyde x:2 FTAL C x F 2x+1 CH 2 CHO C y Perfluorinated aldehyde C y PFAL C y 1 F 2y 1 CHO N Methyl fluorobutane sulfonamide NMeFBSA CF 3 (CF 2 ) 3 SO 2 NHCH 3 N Methyl fluorobutane sulfonamidoethanol NMeFBSE CF 3 (CF 2 ) 3 SO 2 N(CH 3 )CH 2 CH 2 OH N Methyl fluorooctane sulfonamide NMeFOSA CF 3 (CF 2 ) 7 SO 2 NHCH 3 N Ethyl fluorooctane sulfonamide NEtFOSA CF 3 (CF 2 ) 7 SO 2 NHCH 2 CH 3 N Methyl fluorooctane sulfonamidoethanol NMeFOSE CF 3 (CF 2 ) 3 SO 2 N(CH 3 )CH 2 CH 2 OH N Ethyl fluorooctane sulfonamidoethanol NEtFOSE CF 3 (CF 2 ) 3 SO 2 N(CH 2 CH 3 )CH 2 CH 2 OH Perfluorobutane sulfonamide PFBSA CF 3 (CF 2 ) 3 SO 2 NH 2 Perfluorooctane sulfonamide PFOSA CF 3 (CF 2 ) 7 SO 2 NH 2 For fluorotelomer compounds, x is typically an even number between 4 and 12. Number between 4 and 12.


45 Table 1 2 . Summary of PFASs Concentrations Detected in Influent and Secondary Treated Wastewater (as ng/L), and in Sewage Sludge (as ng/g) in Sewage Treatment Plants worldwide. Adapted from Arvaniti and Stasinakis ( 2015 ) . Location PFASs C 4 C 5 C 6 C 7 C 8 C 10 USA and Canada Influent W astewater a (ng/L) (Kentucky, USA) 2.6 6.1 7.0 16 Secondary Treated Wastewater a (ng/L) (Kentucky, USA) 6.3 9.5 8.0 28 Sewage Sludge (ng/g dw) (Kentucky, USA) <2.5 8.2 110 Influent Wastewater a (ng/L) (Georgia, USA) <0.5 4.6 2.5 7.9 Sewage Sludge (ng/g dw) (Georgia, USA) <2.5 38 77 Secondary Treated Wastewater a (ng/L) (New York, USA) <2.5 39 4.0 68 Sewage Sludge (ng/g dw) (New York, USA) <10 18 <10 65 Secondary Treated Wastewater a,b (ng/L) (California, USA) 6.5 24 20 190 Europe Influent W astewater a (ng/L) (Switzerland) 0.6 7.3 0.5 54 18 449 n.d. 160 Influent W astewater a,b (ng/L) (Spain) 19.1 41.9 8.83 78.1 n.d. Secondary T reated W astewater a (ng/L) (Switzerland) n.d. 7.2 2.8 88 16 303 n.d. 4.5 Secondary T reated W astewater a,b (ng/L) (Spain) 57.9 37.7 2.91 91.0 n.d. Sewage S ludge (ng/g dw) (Spain) 0.78 0.01 1.98 41.4 2.15 Sewage S ludge (ng/g dw) (Switzerland) 0.1 28 0.1 6.9 4.0 2440 Asia Influent W astewater a (ng/L) (Korea) 7.4±6.0 7.3±7.3 9.0±13 <0.7 Influent W astewater b (ng/L) (Korea) 110±150 98±180 110±220 <0.7 Influent W astewater a,b (ng/L) (Korea) 37±45 8.8±10 89±150 <0.7 Secondary T reated W astewater a,b (ng/L) (Korea) 30±36 6.7±4.1 110±220 <0.7 Secondary T reated W astewater b (ng/L) (Korea) 140±170 69±120 82±150 0.8±1.0 Secondary T reated W astewater a (ng/L) (Korea) 6.1±3.5 5.0±3.9 6.3±5.0 <0.7 Sewage S ludge (ng/g) (Korea) <1.1 <1.1 15±7.2 <1.5 Sewage S ludge (ng/g) (Korea) 2.1±2.6 6.8±10 260±410 2.4±2.9 Sewage Sludge (ng/g) (Korea) <1.1 <1.1 64±61 1.3±1.7


46 Table 1 2. Continued Location PFASs C 4 C 5 C 6 C 7 C 8 C 10 Australia Secondary T reated W astewater a,b (ng/L) (Australia) n.q. 1.5 2.1 2.2 5.0 Secondary T reated W astewater a,b (ng/L) (Australia) 2.4 6.4 12 36 23 28.6 n.d.: not detected , n.q.: not quantified , a: Municipal Sewage Treatment Plant, b: Industrial Sewage Treatment Plant. n.q.: not quantified. : not analyzed.


47 Table 1 3 . Concentration (ng/L ± 95%CI) of Fluorochemical Analytes in Leachate from Six Landfill Leachates (A D) and a Laboratory Bioreactor . Adapted from Huset et al. ( 2011) . Analyte Site A Site B Site C Site D2 Site D3 Site D6 Laboratory bioreactor PFBA 1700 ± 63 170 ± 6 1400 ± 25 430 ± 34 250 ± 29 540 ± 48 63 ± 22 PFPA 1100 ± 170 120 ± 13 1500 ± 36 730 ± 36 500 ± 29 470 ± 34 460 ± 23 PFHxA 790 ± 50 270 ± 17 620 ± 14 360 ± 12 350 ± 21 430 ± 19 2200 ± 140 PFHpA 328 ± 21 100 ± 14 340 ± 15 170 ± 4.3 150 ± 10 170 ± 3.6 2800 ± 89 PFOA 490 ± 8 1000 ± 19 900 ± 10 380 ± 5.1 490 ± 31 720 ± 60 1100 ± 35 PFNA 23 ± 1.1 22 ± 4.1 28 ± 9.6 20 ± 2.1 19 ± 1.2 26 ± 3.1 140 ± 13 PFDA 15 ± 0.8 14 ± 1.9 23 ± 11 ± 11 0.3 ± 0.8 11 ± 0.5 18 ± 1.4 64 ± 3.7 PFUnDA 0.4 ± 0.6 0 0.1 ± 0.3 0 9.5 ± 1.4 0.9 ± 2.5 0 PADoDA 0.2 ± 0.7 6 ± 1.2 0.8 ± 0.4 0 0.7 ± 1.4 0.2 ± 0.7 8.7 ± 4.4 PFTrDA 0 0.4 ± 0.8 3 ± 1.7 0.2 ± 1.2 18 ± 2 0.7 ± 2.8 5 ± 10 PFTDA 0 1.2 ± 0.9 9 ± 6 2 ± 3 0.7 ± 1.7 13 ± 2.7 10 ± 20 FOUEA 1.5 ± 0.6 10 ± 1.2 0 1.1 ± 1.2 21 ± 2.2 3.2 ± 3 0 PFBS 750 ± 50 280 ± 13 810 ± 36 280 ± 12 390 ± 6.3 890 ± 100 2300 ± 130 PFHxS 700 ± 19 160 ± 8.2 430 ± 13 170 ± 7 200 ± 24 360 ± 110 120 ± 14 PFOS 160 ± 8.6 110 ± 7.5 97 ± 9.2 56 ± 2.5 91 ± 9.9 140 ± 8.9 104 ± 5 PFDS 5.3 ± 1.5 1.1 ± 0.7 0 0.8 ± 0.9 0 1.3 ± 1.2 16 ± 1.6 6:2 FtS 280 ± 11 370 ± 20 280 ± 6.8 29 ± 0.6 56 ± 13 270 ± 67 260 ± 21 8:2 FtS 30 ± 4 120 ± 12 70 ± 7.9 11 ± 1.6 26 ± 4.3 25 ± 1.8 210 ± 25 Me FBSA 1.9 ± 2.8 2.5 ± 17 3.2 ± 3.5 0 0.5 ± 1.6 2.4 ± 2.5 4.2 ± 4.7 Me FBSAA 440 ± 25 79 ± 11 440 ± 33 110 ± 12 58 ± 12 200 ± 14 810 ± 88 FOSA 1.3 ± 1.0 6.6 ± 0.2 0.2 ± 1.4 0 1.4 ± 1.5 0.5 ± 0.8 2.6 ± 1.9 FOSAA 0.7 ± 1.1 1.1 ± 1.2 0.2 ± 1.5 0 0.9 ±1.9 0 12 ± 1.6 Me FOSAA 110 ± 5 280 ± 14 290 ± 19 16 ± 0.4 23 ± 4.7 173 ± 7.1 43 ± 11 Et FOSAA 47 ± 5 480 ± 19 170 ± 24 38 ± 3.5 21 ± 0.7 140 ± 2.4 230 ± 11


48 Figure 1 1 . Timeline of the production, commercialization and legislation of perfluoroalkyl carboxylic acids (PFCAs; at the top) and perfluoroalkyl sulfonic acids (PFSAs; at the bottom). Adapted from Hamid et al. (2018).


49 Figure 1 2 . Typical average concentrations of perfluorooctane sulfonic acid and perfluorooctanoic acid in the blood (serum/plasma) from various countries. Adapted from Jensen and Leffers (2008).


50 Figure 1 3 . Persistent organic pollutants in blood plasm a from pregnant women living in the Norwegian and Russian Arctic. Adapted from Jensen and Leffers (2008).


51 Figure 1 4 . A brief history of PFAS production and regulation. Adapted from Kraft and Riess (2015). C8 Health Project (2005 2013) Environmental awareness & responses from industry PFAS surfactants & polymers to on the market 1950 2000 2005 2010 2015 2020 PFOA found in production plant workers PFASs ubiquitous, bioaccumulating PFASs in th e background population 3M phases out of PFOS related compounds Short F chain alternatives, F eithers Asian production of PFASs ramps up Regulation Europe enforces REACH regulations (2008) PFOS in Annex B of Stockholm Convention (2009) C11 C14 PFCAs vPvB pollutants listed SVHC PFOA in Annex B of Stockholm Convention Stewardship Program (2006 15) eliminates long F chain PFASs


52 C HAPTER 2 PLASMA BASED WATER TREATMENT FOR THE DEGRADATION OF PERFLUOROALKYL SUBSTANCES IN CONTAMINATED WATERS 2.1 Introduction The per and poly fluoroalkyl substances (PFASs) are persistent, bio accumulative, and toxic chemicals (González Barreiro et al. 2006) , which have been used in a wide range of applications including the manufacturing of AFFFs, and several household products such as carpets, paper, and non stick cookware (Fujii et al. 2007, Jahnke and Berger 2009, Prevedouros et al. 2006, Zareitalabad et al. 2013, Lindstrom, Strynar and Libelo 2011, Houtz et al. 2013) . PFAS have been produced only since the 1950s, but because of their extreme resistance to both thermal and biological breakdown as well as their solubility in water (Burns et al. 2008, Rayne and Forest 2009, Liou et al. 2010) , they are pro ne to long distance dispersion . With regard to water pollution, they are detected in relatively high concentrations in effluents of municipal, industrial, and military wastewater treatment plants (Schultz et al. 2006b, Hu et al. 2016) , which in turn act as vehicles for PFASs introduction to natural waterways. Numerous subsets of PFAS with various chain lengths and branched chemical groups have contaminated the environment over time, and perfluoroalkyl acids (PFAAs) consisting of perfluoroalkyl sulfonate (PFSA) and perfluoroalkyl carboxylate (PFCA), are probably the most studied in environmental systems (Rahman, Peldszus, & Anderson,


53 2013) . Perfluoroalkyl sulfonamides (FASAs) an d telomer alcohols (FTOHs) on the other hand drew increasing attention because of the formation of PFAAs as secondary products associated with their degradation (Stock & Furdui, 2007; Xia et al., 2006) . In addition, a wide v ariety of proprietary AFFFs formulations manufactured and sold over the past decades have found their way to the environment as well, making it difficult to identify specific signatures of PFASs in impacted aquatic systems. In fact, depending on the formul ations of AFFFs released to the environment, impacted waterways may contain perfluoroalkyl sulfonates, perfluoroalkyl carboxylates, perfluoroalkyl sulfonamide amino carboxylates, perfluoroalkyl sulfonamido amines, fluorotelomer thio amido sulfonates, fluor otelomer thio hydroxyl ammonium, fluorotelomer sulfonamide betaines to name a few (Backe, Day and Field 2013) . Owing to this complexity, it i s rather challenging to fully understand the fate and behavior of PFASs in the environment (Ra hman et al., 2013) . Human exposure to PFASs includes dietary sources, household dust, air, and of serious concern is drinking water as the main The attention paid to water is in part due to the high aqueous solubility of these compound s, and from the fact that even at relatively low concentrations, PFASs present in water can lead to elevated exposures in the general population (Hu et al. 2016) . PFASs have been detected in ppt levels in drinking water ( Boulanger et al., 2005 ) , and in human tissues (Bartell et


54 al., 2010) . PFAS can lead to adverse he alth effects such as cancer, elevated cholesterol, obesity, suppression of immune system, and endocrine disruption (Barry, Winquist, & Steenland, 2013; Braun et al., 2016; Grandjean & Andersen, 2012) . So far, m ost studies on fate, transport, monitoring and biological impacts of PFAS have focused on perfluoroalkyl sulfonates (C n F 2n+1 SO 3 ) and perfluorocarboxylic acids (C n F 2n+1 COOH); and of special interest are PFOS and PFOA, probably due to the likelihood of human exposure to these compounds. Accordingly, recent regulatory actions have focused on PFOA and PFOS, and a maximum level of 70 ng/L (ppt) for these compounds has been recommended as lifetime health advisory for drinking water by the USEPA (USEPA, 2016) . The above health concerns and the action limit of 70 ppt have stimulated research for advanced treatment processes to remove PFAS from contaminated waters. Sev eral treatment techniques have been proposed for both ex situ and in situ scenarios. The removal of PFAS from contaminated waters using traditional water treatment techniques (e.g. ferric or alum coagulation, granular/micro /ultra filtration, aeration, ox idation, and disinfection) is mostly ineffective (Appleman et al., 2013) . In contrast, anion exchange and granular activated carbon technologies preferably removed longer chain PFAS, while reverse osmosis demonstrated significant removal efficiencies for both short and long chains compounds (USEPA, 2016; Mccleaf et al., 2017) . However, these techniques rely on pollutant phase transfer, which has


55 implications for the disposal of PFAS transferred onto solid sorbents. Other reported issues associated with phase transfer methods include the desorption of previously sorbed short chain PFAS during co removal studies of mixtures of long and short chain compounds (Mccleaf et al., 2017) . In addition to the diffe rent techniques based on phase transfer, degradation methods have also been proposed and they include AOPs, microbial biodegradation, and sonochemical destruction (Kucharzyk, Darlington, Benotti, Deeb, & Hawley, 2017) . Results associated with these techniques suffer from a few limitations such as slow and incom plete de fluorination (biodegradation), poor degradation efficiency and high cost (for sonochemistry), and lack of knowledge on the potential degradation products in the case of AOP based treatments (Kucharzyk, Darlington, Benotti, Deeb, & Hawley, 2017) . Plasma based degradation technologies have also been prop osed for treatment of PFASs contaminated waters (Stratton et al. 2017, Yasuoka et al. 2010) . Similar to AOPs, this approach makes use of highly reactive species, but which are produced in situ, often with little to no chemical inputs. However, most of the previous attempts to degrade PFA S by this approach involved the use of rather inefficient reactor types and DC discharges, with limited successes. Recently, and using pulsed electrical discharges instead of DC discharges, Stratton et al. achieved higher PFOA removal and better de fluorin ation efficiencies (Stratton et al. 2017) . In this stud y, we used PFOA and PFOS contaminated water samples with


56 different chemical compositions to evaluate the potential of a Dielectric Barrier Discharge (DBD) plasma based technique to improve the removal efficiencies using a reactor that consumes an average p ower of ~10 watts . Conducted laboratory experiments were designed to ascertain both the efficiency and the adequacy of this proposed advanced water treatment technique through degradation and characterization of fragments resulting from the degradation pro cess. 2.2 Material and M ethods 2.2.1 Water Samples Collection To determine the efficiency of the plasma based water treatment, PFOA and PFOS were used as example PFASs to spike waters with different characteristics including deionized water, treated municipal wastewater effluents, and landfill leachate. Treated waste water effluent samples were obtained from three different municipal WWTPs in the city of Gainesville, Florida. A landfill leachate sample was obtained from New River Regional Landfill (NNRL) in Union County, Florida. The initial dissolved organic concentra specific UV 254 absorbance (SUVA 254 ) of 3.8 L/mg·m (with ), implying a DOC mixture from both terrestrial and microbial origins (Oppong Anane et al. 2018) . It was a colored fluid and only diluted solutions (1%, v/v in deionized water) were used in this study. All liquid samples were collected, processed in prewashed glass bottles and refrigerated (4 o C) pending analysis.


57 2.2.2 Reagents, Standards and Charac terization of the Tested Water Samples High purity PFOA and PFOS were purchased from Sigma Aldrich (St. Louis, MO) and Santa Cruz Biotechnology Inc. (Dallas, TX), respectively. Standard stock solutions for these compounds were prepared separately by dissol ving a known amount of each of the compounds into a known final volume of HPLC grade methanol. For the extraction of PFOA and PFOS from actual water samples, which was then followed mass spectrometry analysis, methylene chloride, 2 propanol for ultra trace analysis, HPLC grade methanol, and trace metal grade nitric acid purchased from Fisher Scientific was used. Standard solutions for cations (Li + , Na + , , K + , Mg 2+ , Mn 2+ , Ca 2+ and Sr 2+ ) and anions (F , Cl , , , Br , and ) analysis by ion chromatography were purchased from Dionex. The water samples (treated wastewater effluents and diluted landfill leachate) used in this study were analyzed for pH, alkalinity and major ions concentrations. The pH was measured using an Accu met® AE150 pH meter from Fisher Scientific. Total alkalinity was determined by titration using 0. 2 N sulfuric acid and pH 4.3 as the endpoint. Dissolved major ions were determined on S 3000 from Dionex. 2.2.3 Water Sample Spiking with PFOA and PFOS Spiked water samples were prepared using the above described PFOA or PFOS stock solutions, and through the addition of known aliquot volumes of these


58 concentrated solutions to waters to pr oduce desired concentrations. Briefly, a given spiked water sample was divided into 6 sub samples. Three of these six sub samples were not treated with plasma and used as a control group. The other 3 underwent the proposed plasma water treatment technique (treated group), and the results of these two groups compared. Tested concentrations of PFOA and PFAS ranged from ppb to ppm, depending on the objective of the different experiments. 2.2.4 Plasma Water Treatment A custom experimental setup to treat the spi ked water samples mentioned above was designed and carried out by SurfPlasma Inc. The motivation behind this experiment was to generate reactive oxygen species (ROS) using an efficient DBD plasma generator (Choudhury et al. 2018) to be injected and mixed into the water sample. The current diffusion method to mix ROS into water was inspired by previous studie s in which an air pump was used with a diffusion stone to transfer the ROS into water. The plasma generator used is based on a DBD capacitive power circuit that operates as a power inverter converting a low AC/DC input voltage into a high frequency (kHz) hig h voltage (~7 kVpp), which is then used to power the plasma reactor electrodes (Portugal, Roy and Lin 2017) . The reactor consumes an average power of ~10 watts, and is quite efficient for generating ROS including ozone.


59 This setup co nsisted of a Dielectric Barrier Discharge (DBD) plasma generator, an impeller driven water pump, a venturi injector, and a static mixer and is outlined in Figure 2 1 . Contaminated water circulates through the setup via the impeller pump at a flow rate of 1 .4 GPM. As the water passes through the venturi injector it is forced through a conical body, initiating a pressure difference between the inlet and outlet ports thereby creating a vacuum inside pulling air through a suction port. The suction port is conne cted to a volume in which the plasma generator produces ROS. After mixing in the venturi injector, the ROS is further mixed into the water with an in line static mixer. Helical structures in the static mixer produce flow division and radial mixing. The wat er is pumped in and out of an open air beaker where any ROS that is not fully transferred into the water escape into the atmosphere. The process was run for 40 minutes in which the water was recirculated and continuously injected with actively generated RO S. 2.2.5 Extraction and Analysis of PFOA and PFOS To extract PFOA and PFOS from water samples, an extraction method adapted from peer reviewed literature was used (Barco et al., 2003) . Briefly, 200 mL of PFOA and PFOS spiked samples from control and plasma treated groups were placed in beakers. The pH of each sample was first adjusted to 2 using a nitric acid solution. The sample was then transferred into an organic solvent extraction separator, and a 100 mL mixed solution of methylene chloride and 2 propanol (9:1, v/v) added. Following a


60 thorough mixing by shaking for 20 minu tes, the mixture was allowed to separate into aqueous and organic phases, and the water fraction discarded. Next, the organic fraction was transferred into a glass container and placed on a heating plate adjusted at 40 °C. The organic solvent was evaporate d to dryness under a gentle nitrogen flow. Finally, 2 mL of methanol was added to the dry container to solubilize the PFOA or the PFOS and used for their analysis at the Mass Spectrometry Research and Education Center (MSREC), on the University of Florida campus. Following the above described sample extraction process, FPOA and PFOS recovered in methanol extracts were first separated from all the other organic compounds potentially present in the treated complex water matrices. To do so, chromatographic sep aration was performed using reverse phase HPLC equipped with a C 8 column as solid phase and water/methanol in a gradient mode as a mobile phase. This first step allowed the identification of the retention time of compounds of interest prior to mass spectrometry analysis. Mass spectra of PFOA and PFOS were obtained by us e of electrospray ionization double mass spectrometry (ESI MS/MS) in either negative or positive mode. This technique produces molecular ions as well as molecular fragments characterized by specific mass to charge ratios (m/z). For example, the ( )ESI MS/M S of the m/z 413 [M H] ion of PFOA efficiently produced an abundant m/z 369 product ion with some additional relatively minor m/z 347, 219 and 169 product ions. The m/z 499 [M H] ion of PFOS only produced a series of low intensity product


61 ions, m/z 169, 230, 280, 330, 380 and 419. Due to the low m/z cutoff of the quadrupole ion trap mass spectrometer, it is not possible to detect the major m/z 80 and m/z 99 product ions of the m/z 499 [M H] ion of PFOS with internal MS/MS. Therefore, source collision ind uced dissociation (SCID) was used to confirm the identity of compounds in the produced spectra , either in source collision induced dissociation (SCID) MS or in trap CID MS/MS mode . The optimum ( )ESI SCID occurred at SCID 10 50V. For semi quantitation of the relative amounts the PFOS and PFOA in treated and untreated water, ( )ESI MS full scan (m/z 200 1200) was used and the peak areas of the sum of their [M H] and [ 2M H] ions were determined. For the detection of the degradation products, HPLC/( )ESI MS and (+)ESI MS screens were initially done followed with selected data dependent MSn analyses on the treated PFOA and PFOS in water samples. 2.3 Results and D iscussio ns 2.3.1 Chemical Characterization of the Different Water Samples Used in this Study Relevant water quality parameters for samples used in this study are given in Table 2 1. Samples WWTP 2 and WWTP 3 from the city of Gainesville are very similar in their c hemical composition, with low ionic strength values of 0.0072M and 0.0076M, respectively. In contrast, the water samples WWTP 1 (which treats wastewater from the


62 ionic s trength, with respective values of 0.0105M and 0.4059M. Concentrations of DOC were not available at the time of this reporting and will be determined later for inclusion in the anticipated manuscript. Overall, these differences in water chemistries could a llow for the assessment of the effects of water chemical composition on the efficiency of plasma based water treatment techniques, if any. 2.3.2 Detection of PFOA and PFOS PFOA and PFOS in non treated and plasma treated water samples were analyzed at the Mass Spectrometer Research Education Center at the University of Florida. Figure 2 2 shows example spectra from an ESI MS used in negative mode, with clear identification of the compounds of interest, when pure PFOA and PFOS compounds are dissolved in met hanol (i.e. standard solutions). For the two compounds, we can see molecular ion peaks at m/z of 412.3 and 499.2 for PFOA (MW 413g/mol) and PFOS (MW 500g/mol), respectively. In addition, and on the far right of the spectra, peaks of combined molecular ions appear at m/z of 848.8 and 1020.8 for PFOA and PFOS, respectively. Finally, while several fragments can be seen in the PFOA spectrum, fragmentation seems to be minimal with PFOS. These results are presented to simply illustrate the ability of the equipmen t used to detect PFOA and PFOS, at concentrations ranging from low ppb to ppm.


63 2.3.3 Effect of Plasma based Treatment Procedure on the Degradation of PFOA and PFOS Figure 2 3 shows the results of PFOA and PFOS degradation experiments using the plasma base d water treatment process at two water recirculation rates of 0.8 and 1.4 GPM. Overall, the two procedures resulted in significant degradation of both compounds. However, treatment 1 (1.4 GPM) had better removal efficiencies for both compounds and regardle ss of the water recirculation rate. It can also be noted that under similar conditions (such as water chemical compositions, similar concentrations of PFOA and PFOS, and treatment conditions and time), PFOA is more prone to plasma induced degradation than PFOS. Finally, the diluted landfill leachate (LL) was processed using only the above described Treatment 1, and again, PFOA removal was nearly complete while the breakdown of PFOS was less pronounced. 2.3.4 Optimization of the Degradation of PFOA and PFOS From the above discussed degradation results, the first treatment (or treatment 1) was retained as the most efficient. It was then used in further laboratory studies to fine tune the performance of the proposed plasma based water treatment to degrade PFOA and PFOS. Selected results from these investigations and using WWTP 1 as example water sample are presented in Figures 2 4 and 2 5. These are chromatograms of PFOA and PFOS in non treated water samples (top graphs in Figures 2 4 and 2 5), and in plasma tre ated water samples (bottom figures). Similar results were obtained with the other


64 water samples, and every time, a residual PFOS was detected (Figure 2 5 bottom) while PFOA was completely degraded (Figure 2 4 bottom). When comparing the performance of this plasma based water treatment process to the results of other PFASs degradation techniques reported earlier in the literature (Figure 2 6), it obvious that the former has better removal efficiencies, 100% for PFOA and 98% for PFOS. This in comparison with some of previous results such as <10% for PFOA and 10 to 50% for PFOS using AOPs (Cummings et al. 2015) , 67% removal of PFOS by microbial degradation (Kwon et al. 2014) , and degradation rates of 46% and 66% for PFOA and PFOS by sonochemical degradation (Cheng et al. 2008) . A previous study using plasma (Stratton et al. 2017) reported degradation rates of up to 90% for PFOA with a power input of 76.5 W, but only 25% degradation rate when the power input was lowered to 4.1 W. Comparatively, this study accomplishes a 100% removal of PFOA with a power input of <10 W, which is a significant improvement. More recently, a study on the removal of PFOA and PFOS from contaminated waters using a plasma based water treatment process and employing argon gas at a flow rate of 4 L/min achieved an efficiency of 90% removal for both of PFOA a nd PFOS with treatment times of 40 and 60 minutes (Singh et al. 2019) . While the effectiveness of plasma based water treatment appears is encouraging overall, it raises questions about the types and fate of potential degradation product s formed during the breakdown of PFASs initially present in the sample. Ideally, one


65 would seek complete mineralization, and therefore, not a production of smaller fluorinated carbon compounds, as they may be as toxic as their parent compounds, or even wor se. The logic step from this point is, therefore, to conduct the identification of the potential degradation products when PFOA and PFOS are submitted to plasma treatment. For instance, depending on the plasma system used (low or high voltage power supply, frequencies of generated pulses, mode of discharge), highly reactive species can be formed, which are responsible for the degradation of PFASs. In fact, in a rather comprehensive study by Singh et al., perfluoroheptanoic acid (PFHpA), perfluorohexanoic ac id (PFHxA), perfluoropentanoic acid (PFPeA), PFBA were found as common byproducts of both PFOA and PFOS degradation, with two additional byproducts (perfluorohexanesulfonate (PFH X S) and perfluorobutanesulfonate (PFBS)) detected for PFOS only (Singh et al. 2019) . However, the same study concluded that the concentrations of the degradation products decreased with the length of sample exposure to plasma, suggesting that these could be just intermediate species towards complete mineralizatio n. 2.3.5 Identification of the Degradation Products of PFOA and PFOS Following Plasma Treatment. In a plasma based water treatment such as the one used in this study, use of electricity converts water into a mixture of highly reactive species such as OH, H , HO 2 , O 2 , H 2 O 2 , and aqueous electrons ( ) called plasma (Singh et al. 2019) . Preliminary


66 work was undertaken in this study to identify the potential degradation products of PFOA, resulting in the following observations. a. P resence of peaks not presented prior to plasma treatment. Figure 2 7 shows an example illustration of potential degradation products, with m/z 272 ion peaks, present in the plasma treated DI water containing either PFOA (Figure 2.6c) or PFOS (Figure 2.6d). This m/z 272 ion peak is absent from chromatograms of DI water untreated (a) or treated (b) with plasma. Similarly, the m/z 272 ion peak is absent from chromatograms of PFOA and PFOS containing waters when untreated with plasma (Figure 2 7e and f). Unfort unately, the mass spectra of this m/z 272 ion peak were not determined, making it difficult to positively confirm the presence of degradants as (CF 2 ) or (CH 2 ) . Pending the identification of this fragment, and based on the conclusions from Singh et al m entioned above (Singh et al. 2019) , extending the treatment time from the current 40 min used in this study to 60 min could eliminate these ion peaks and therefore these fragments. b. The increase in the peak areas of peaks detected pri or to plasma treatment. In contrast to the new ion peaks discussed above, those detected by (+)ESI MS with m/z of 328.7 in plasma treated plain water (Figure 2 8 b) and plasma treated water containing PFOA and PFOS (Figures 2 8c and 2 8d) show simply incre ased peak areas. The ion peaks in the plasma treated PFOA and PFOS containing samples are ~22 times higher than the same ion peak in the treated water sample (see inset table in


67 Figure 2 8). Without knowledge of the structure of this compound, one can only speculate that these peaks could be related to degradation products (assuming the presence of PFOA and PFOS contaminants in DI water). On the other hand, these peaks can be due to the concentrations of degradants from other materials than PFOA and PFOS pr esent in the water, but with PFOA and PFOS acting as catalysts. Again, in the absence of mass spectra for structure determination, extending the time of plasma treatment time could help explain these ion peaks as they either increase or disappear over time . Overall, the above results point to a successful degradability of the tested PFASs using a plasma based water treatment process. However, research is needed to confirm the complete defluorination, and therefore, the lack of fluorinated daughter products such as FAAs.


68 Table 2 1 . Chemical characterization of the different water samples tested. Samples labeled WWTPs were collected from wastewater treatments plants in 3 different locations in the city of Gainesville, Florida. Water quality parameters Water samples tested in this study UF: Wastewater effluent (WWTP 1) TREOO: Wastewater effluent (WWTP 2) MS: Wastewater effluent (WWTP 3) Landfill: Diluted landfill leachate (LL) pH 6.84 6.93 6.61 7.59 Alkalinity (mg CaCO 3 /L) 66 82.70 125 5,409 Cations (mg/L) Li + 0.025 0.003 0.008 n.d. Na + 56.91 47.32 58.70 1.44 2.29 2.40 n.d. 1.70 K + 11.523 9.78 5.20 0.49 Mg 2+ 28.05 18.37 11.95 0.12 Sr 2+ 46.93 35.43 50.44 4.02 Anions (mg/L) F 0.32 0.23 0.12 n.d. Cl 76.16 71.06 89.21 1.71 5.23 5.30 6.77 0.16 188.16 93.71 76.82 0.31 Br n.d. 0.7361 n.d. n.d. 10.84 14.72 13.45 0.24 n.d. n.d. 0.48 n.d. **Ionic strength (M) 0.0105 0.0072 0.0076 0.4059 n.d.: not detected, UF: University of Florida WWTP, MS: Gainesville Main Street WWTP (residential and industrial), TREOO WWTP located in the West side of the city of Gainesville (primarily a residential area). **The ionic strength includes the alkalinity as bicarbonate ion.


69 Figure 2 1 . The process of ROS Water Treatment System.


70 Figure 2 2 . Illustration of example spectra determined using sample WWTP 2 and analysis by the negative electrospray ionization mass spectrometry (abbreviated ( )ESI MS). PFOA (MW 414 g/mol) produced a m/z of 413 [M H] and m/z 849 [2M H] ions (top spectrum). Similarly, PFOS (MW 500g/mol) produced a m/z of 499 [M H] and m/z 1021 [2M H] ions (bottom spectrum). In each case, the sum of two ions can be used to estimate the relative amounts present. P erfluorooctanoic acid (PFOA) Perfluorooctane sulfonate (PFOS)


71 Figure 2 3 . Effects of two water treatment procedures on the removal efficiency of PFOA and PFOS. Treatment 1corresponds to plasma application, ozone generation and water recirculated at a rate of 1.4 GPM for 40 min. Treatment 2 is similar to Treatment 1, but at a much lower recirculation rate of 8.0 GPM.


72 Figure 2 4 . Chromatograms of the degradation PFOA using WWTP 1 water as an example, with the plasma non treated (top) and treated (bottom) shown comparatively. Results were obtained using an HPLC/( )ESI MS mass chromatograms of the sum of the m/z 413 [M H] and the m/z 849 [2M 2H+Na] . PFOA in WWTP 1 Untreated PFOA in WWTP 1 Plasma T reated


73 Figure 2 5 . Chromatograms of the degradation PFOS using WWTP 1 water as an example, with the plasma non treated (top) and treated (bottom) HPLC/( )ESI MS mass chromatograms of the sum of the m/z 499 [M H] and the m/z 1021 [2M 2H+Na] . PFOS in WWTP 1 Untreated PFOS in WWTP 1 Plasma T reated


74 Figure 2 6 . Co mparative performances of different PFASs degradation techniques based on removal efficiencies published in the literature and data from this study.


75 Figure 2 7 . Potential degradation products with m/z 272 ion peaks present in the plasma treated DI water containing either PFOA (c) or PFOS (d). (a) ( b ) ( c ) ( d ) (e) ( f )


76 Figure 2 8 . Chromatograms showing the (+)ESI MS ion peak with an m/z of 328.7 detected in plasma treated water (b) and in the plasma treated water samples containing PFOA and PFOS. The insert Table shows the increase in the areas of the ion peaks by >22 times than in the treated water sample. (a) (b) (c) (d) (e) (f)


77 CHAPTER 3 SUMMARY AND C ONCLUSIONS 3.1 Summary of Research This research was designed to test the feasibility and the removal efficiency of selected PFAS compounds from contaminated waters, based on different water recirculating flow rates and plasma based treatment . The results indicated that the different flow rates used during treatment conducted at room temperature contribute to the different removal efficiencies for PFOA and PFOS. The optimum removal efficiencies measured in this study were 92.5% and 100% for PFOS and PFOA, respectively. To identify the chemical structures of byproduct s from PFOS and PFOA in water, treated deionized water with equivalent concentrations of PFOS and PFOA by the same technique with the higher removal efficiency was used to analyze in order to avoid the impurity. Preliminary investigation into the byproduct s showed there were peaks not present prior to plasma treatment, and a peak with increased peak areas compared with untreated DI water. Although non identification of particular breakdown fragments from PFOA and PFOS was obtained, the study confirmed the f easibility of the new plasma based treatment with high removal efficiency for PFASs in water. 3.2 Future Work Recommendations From the results presented in this thesis , it is obvious that the plasma based technique is efficient at removing both PFOA and P FOS from contaminated aqueous


78 solutions . Studies on the identification of the potential fragment s produced by the degradation process were less conclusive and should be the focus of future work. The latter limits the statement of solid conclusions on the p otential mechanisms of FFOS and PFOA breakdown under degradation conditions used in this research. Additionally, further studies are needed to determine the role of key water parameters such as pH, time of exposure to plasma, and types of reactive oxygen s pecies on the degradation of PFAS. It is important to note that, unlike most treatment techniques, which are currently available, a successful plasma based technique will have no byproducts , which is a tremendous advantage.


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96 BIOGRAPHICAL SKETCH Meiting Song was born in 1994, in Luoyang, Henan Province, China . She graduated with a bachelor's degree in e n vironmental s cience from Henan Agricultural University in Zhengzhou , China 2017. She joined the University of Florida in fall 2017 , and after four semesters of study she graduate d with a m ast e d egree . Her research focused on the treatment of water contaminated with emerging po llutants.