ABSTRACT. Perfluorinated compounds (PFCs) are surfactants used in the manufacturing of numerous

Transcription

1 ABSTRACT DUDLEY, LEIGH-ANN MARIE BENDER. Removal of Perfluorinated Compounds by Powdered Activated Carbon, Superfine Powdered Activated Carbon, and Anion Exchange Resins. (Under the direction of Detlef R.U. Knappe). Perfluorinated compounds (PFCs) are surfactants used in the manufacturing of numerous consumer products, including stain repellents, nonstick coatings, and water repellent fabrics. Because of their persistence, PFCs can be detected globally in aquatic environments, including drinking water sources. In 2009, the EPA issued a drinking water Provisional Health Advisory for perfluorooctanoic acid (PFOA) and perfluorooctane sulfonate (PFOS) at 0.4 and 0.2 μg/l, respectively, to protect consumers against adverse health effects resulting from short-term PFOA and PFOS exposure. In New Jersey, a health-based PFOA guideline of 0.04 μg/l was developed as a protective measure against long-term PFOA exposure. PFOA levels of 0.04 μg/l are not uncommon in drinking water sources. To date, little is known about the effectiveness of drinking water treatment processes for the removal of PFCs at drinking water relevant concentrations. The objective of this research was to assess the effectiveness of powdered activated carbon (PAC) adsorption and anion exchange processes for the removal of PFCs from drinking water sources. To gain insights into the behavior of PFCs in activated carbon and ion exchange treatment processes, experiments were conducted with a homologous series of 7 perfluorocarboxylic acids [perfluorobutanoic acid (C4) to perfluorodecanoic acid (C10)] and 3 perfluorosulfonic acids [perfluorobutane sulfonate (PFBS), perfluorohexane sulfonate (PFHS), PFOS]. PFC uptake rates were compared for five PACs, prepared from different base materials, and five

2 superfine PACs (S-PACs), prepared by wet-milling as-received PACs. Three strong base anion (SBA) exchange resins and one weak base anion (WBA) exchange resin were evaluated for PFC removal. Experiments were conducted in buffered ultrapure water and two drinking water sources. In buffered ultrapure water, PFC adsorbability increased with increasing PFC carbon chain length: C4 removal was negligible at a (S-)PAC dose of 15 mg/l while C10 removal was almost complete. For PFCs of a given carbon chain length, sulfonates were more readily removed than carboxylates. Among the tested carbons, thermally activated wood and coconut shell-based S-PACs were the most effective for PFC removal: for eight of ten PFCs, >85% removal was achieved from buffered ultrapure water (adsorbent dose: 15 mg/l). The presence of natural organic matter (NOM) reduced the PFC removal effectiveness of (S-)PACs. In North Carolina (NC) reservoir water with a total organic carbon (TOC) concentration of 4.5 mg/l, >85% removal could be achieved for only three of the ten PFCs (perfluorononanoic acid, C10, PFOS) with the coconut shell based S-PAC. Among the asreceived PACs PFC removal was highest at 55% for C10 and PFOS when 15 mg/l thermally activated wood-based PAC was added to NC reservoir water. PFC uptake by anion exchange resins also increased with increasing PFC carbon-chain length, and sulfonates were more readily removed than carboxylates. Polystyrene-based SBA resins achieved the highest PFC removals. For C4, the most challenging PFC to remove, 90% removal from buffered ultrapure water was achieved with a resin dose of 5 ml/l. While PFC uptake was slightly lower with the polyacrylic SBA resin, uptake kinetics were faster. The

3 WBA resin could not achieve more than 60% PFC removal from buffered ultrapure water at the highest tested resin dose of 10 ml/l. Compared to buffered ultrapure water, removal of all 10 PFCs increased in the presence of NOM. The most likely explanation is that NOM altered the resin surface such that its affinity for PFCs increased. The addition of 200 mg/l sulfate, an anion with high affinity for anion exchange resins, to NC reservoir water decreased PFC removal. Overall, the results of this research show that adsorptive removal of C4, perfluoropentanoic acid, perfluorohexanoic acid, perfluoroheptanoic acid, and PFBS from drinking water sources requires activated carbon doses that are too high to be practical. In contrast, polystyrene-based SBA exchange resins effectively removed all 10 PFCs from drinking water sources. Before the application of anion exchange for PFC removal can be recommended, effective resin regeneration strategies need to be developed. Alternatively, a hybrid adsorption/anion exchange treatment strategy may be most effective for PFC removal, in which more strongly adsorbing PFCs are initially removed by activated carbon and the more weakly adsorbing PFCs subsequently by anion exchange.

4 Removal of Perfluorinated Compounds by Powdered Activated Carbon, Superfine Powdered Activated Carbon, and Anion Exchange Resins by Leigh-Ann Marie Bender Dudley A thesis submitted to the Graduate Faculty of North Carolina State University in partial fulfillment of the requirements for the Degree of Master of Science Environmental Engineering Raleigh, North Carolina 2012 APPROVED BY: Joel J. Ducoste Francis L. de los Reyes Detlef R.U. Knappe Chair of Advisory Committee

5 DEDICATION To my parents, my sister, Tina, my brother, Kathan, and my loving husband, Harrison for all of their love and support. ii

6 BIOGRAPHY Leigh-Ann Dudley graduated from North Carolina State University in May 2010 with a Bachelor of Science in Chemical Engineering. She continued her studies toward a Master of Science in Environmental Engineering at North Carolina State University under the direction of Detlef R.U. Knappe. iii

7 ACKNOWLEDGMENTS I would like to thank the following people and organizations for their participation and support of this project: Water Research Foundation for funding this project (Project #4344) Detlef Knappe for his support and guidance throughout this project, undergraduate, and graduate studies Mark Strynar, Andrew Lindstrom, and Larry McMillan of the US Environmental Protection agency for opening up their lab and sharing their expertise throughout this project Joel Ducoste and Francis de los Reyes for serving on my committee and being supportive mentors throughout graduate classes Miguel Arias with Orica for providing expertise and donating resin samples Purolite for donating resin samples Calgon Carbon Corporation, Jacobi Carbons, MeadWestvaco Corporation, Norit Americas, and Pica USA, Inc. for donating powdered activated carbons (PACs) John Hill and Harry Way at NETZSCH Premier Technologies, LLC. For grinding asreceived PAC into superfine powdered activated carbons Orange Water and Sewer Authority and Cincinnati Water Authority for providing water David Black, Environmental Lab Manager, for his support iv

8 My research group and friends including Bilgen, Qianru, Angela, Susan, Meredith, Allison, Yuxuan, Kara and Brandon for their support and friendship Everyone in the Broughton Environmental Lab and Mann 319A office for making this a wonderful place to work and learn v

9 TABLE OF CONTENTS LIST OF TABLES... ix LIST OF FIGURES... xi ABBREVIATIONS... xiv CHAPTER 1: INTRODUCTION AND OBJECTIVES...1 MOTIVATION... 1 RESEARCH OBJECTIVES... 3 APPROACH... 4 CHAPTER 2: BACKGROUND...6 PERFLUORINATED COMPOUNDS... 6 Characteristics... 6 Occurrence... 9 Health Effects Health Advisories Effectiveness of Conventional Water Treatment Processes for PFC Removal ACTIVATED CARBON ADSORPTION Powdered Activated Carbon Effectiveness of Activated Carbon Adsorption for PFC Removal ION EXCHANGE Anion Exchange Resins Effectiveness of Anion Exchange for PFC Removal SUMMARY vi

10 CHAPTER 3: MATERIALS AND METHODS...23 MATERIALS Water Activated carbons Anion Exchange Resins Perfluorinated Compounds METHODS (S-)PAC Batch kinetic tests (S-)PAC Isotherm tests Isotherm Modeling Approach (S-)PAC Jar Tests Batch Kinetic Tests Conducted with Anion Exchange Resin PFC analysis Total Organic Carbon Analysis CHAPTER 4: RESULTS AND DISCUSSION...37 PFC REMOVAL BY ACTIVATED CARBON (S-)PAC Physical and Chemical Characteristics (S-)PAC Batch Kinetic Tests (S-)PAC Isotherms (S-) PAC Jar Tests (S-)PAC Dose Requirements for PFC Removal PFC REMOVAL BY ANION EXCHANGE CHAPTER 5: SUMMARY AND CONCLUSIONS PFC Removal with Activated Carbon PFC Removal with Anion Exchange Resins REFERENCES vii

11 APPENDICES Appendix A: Evaluation of PFC Loss to Bottles and Jars Appendix B: Representative LC-MS/MS Chromatograms Appendix C: Representative Calibration Curves from PFC Analysis Appendix D: Freundlich Exponent and Initial Concentration for EBC Appendix E: Comparison of Anion Exchange Resin Performance in Three Waters viii

12 LIST OF TABLES Table 2.1 Chemical properties of PFCs studied... 8 Table 2.2 Summary of studies detecting PFCs in water Table 3.1 Water quality parameters for drinking water sources Table 3.2 As-received PACs Table 3.3 S-PAC Inventory Table 3.4 Anion exchange resin properties Table 3.5 Perfluorinated compounds selected for this study Table 3.6 Jar test mixing regime Table 3.7 LC Gradient Method for PFC Analysis Table 3.8 Mass transitions for PFCs and internal standards Table 4.1 Surface area and pore volume analysis Table 4.2 Elemental analysis Table 4.3 Freundlich K and 1/n values for carbons C and S-C Table 4.4 Equilibrium solid phase concentrations calculated for equilibrium concentrations of 10 and 100 ng/l Table 4.5 Carbon Doses Required to Achieve 90% PFC Removal from CCR and ORW Table 4.6 Carbon Dose Required to Achieve 50% Removal of Table 4.7 Carbon Doses Required to Achieve 90% PFC Removal from CCR in Isotherm and Jar Tests Table 4.8 Resin Dose Requirements for 90% PFC Removal from CCR ix

13 Table D.1 Freundlich parameters in CCR and ORW with C.157 Table D.2 Freundlich parameters in CCR and ORW with S-C x

14 LIST OF FIGURES Figure 2.1 General chemical formulas for (a) perfluorocarboxylic acids (b) perfluorosulfonic acids Figure 2.2 Strong-base anion exchange resin (Source: Clifford, Sorg, and Ghurye, 2011) Figure 4.1 Effect of (S-)PAC type and particle size on PFC adsorption kinetics Figure 4.2 Comparison of C7 uptake rates from single-solute solution and PFC mixture Figure 4.3 Effect of activated carbon type and particle size on PFC uptake rates from CCR.47 Figure 4.4 Effect of ph and ionic strength on PFC removal by (S-)C Figure 4.5 Effect of ph, hardness, and turbidity in natural waters on PFC removal Figure 4.6 Comparison of UV 254 removal in CCR and ORW. Carbon Dose: 15 mg/l. Contact Time: 2 hours Figure 4.7 PFC adsorption isotherms obtained with as-received carbon C in UPW: (a) carboxylates, (b) sulfonates Figure 4.8 PFC adsorption isotherms obtained with superfine carbon S-C in UPW: (a) carboxylates, (b) sulfonates Figure 4.9 PFC adsorption isotherms obtained with as-received carbon C in CCR: (a) carboxylates, (b) sulfonates Figure 4.10 PFC adsorption isotherms obtained with superfine carbon S-C in CCR: (a) carboxylates, (b) sulfonates Figure 4.11 PFC adsorption isotherms obtained with as-received carbon C in ORW: (a) carboxylates, (b) sulfonates xi

15 Figure 4.12 PFC adsorption isotherms obtained with superfine carbon S-C in ORW: (a) carboxylates, (b) sulfonates Figure 4.13 Comparison of PFC adsorption isotherms for (S-)PACs in UPW, CCR, and ORW Figure 4.14 Effect of (S-)PAC type and dose on PFC removal in jar tests Figure 4.15 Comparison of UV 254 removals for different (S-)PAC types and doses Figure 4.16 Effect of carbon type and dose on settled water turbidity. Alum dose: 55 mg/l. Coagulation ph: Figure 4.17 Effect of coagulation ph on PFC removal. Alum dose: 55 mg/l. (S-)C dose: 15 mg/l Figure 4.18 Effect of coagulation ph on UV 254 removal. Alum dose: 55 mg/l. (S-)C dose: 15 mg/l Figure 4.19 Comparison of the effect of coagulation ph on settled water turbidity Figure 4.20 Effect of the timing of (S-)PAC addition on PFC removal. Alum dose: 55 mg/l. Coagulation ph: 6.2. (S-)C dose: 15 mg/l Figure 4.21 Effect of the timing of (S-)PAC addition on settled water UV 254 absorbance. Alum dose: 15 mg/l. Coagulation ph: 6.2. (S-)C dose: 15 mg/l Figure 4.22 Effect of the timing of addition of coagulant and (S-)PAC on settled water turbidity Figure 4.23 Effect of (S-)PAC dose on PFC removal from CCR at equilibrium (isotherm results) and non-equilibrium (jar test results) conditions xii

16 Figure 4.24 Effect of (S-)PAC dose on PFC removal from ORW at equilibrium (isotherm results) and non-equilibrium (jar test results) conditions Figure 4.25 PFC uptake kinetics from amended UPW for four anion exchange resins Figure 4.26 Effectiveness of four anion exchange resins for PFC removal from two drinking water sources (CCR and ORW) Figure 4.27 UV 254 Removal with resins 1-4 in CCR Figure 4.28 UV 254 Removal with resins 1-4 in ORW Figure 4.29 Effect of resin contact time on PFC removal from CCR and ORW. Resin dose: 5 ml/l Figure 4.30 Effect of background water matrix on PFC removal with resin Figure 4.31 Comparison of PFC removals obtained with resin 1 following NOM preloading and with original resin 1. Water: CCR Figure A.1 Comparison of time 0 and 2 and 3 week PFC concentrations in isotherm bottles.146 Figure A.2 Comparison of time 0 and 46 minutes PFC concentrations in acrylic jars 146 Figure B.1 Chromatograms for carboxylates C4-C Figure B.2 Chromatograms for sulfonates PFBS, PFHS, and PFOS Figure B.3 Chromatograms for internal standards Figure C.1 Calibration curves for ten PFCs Figure E.1 Comparison of PFC removal achieved in three waters with resin Figure E.2 Comparison of PFC removal achieved in three waters with resin Figure E.3 Comparison of PFC removal achieved in three waters with resin xiii

17 ABBREVIATIONS BET C4 C5 C6 C7 C8 C9 C10 CaCl 2 CCR DI DVB EBC EPA HCl hr IAST IS IS L LC-MS/MS LOAEL LOQ mg ml mm mm min M Brunauer-Emmett-Teller perfluorobutanoic acid perfluoropentanoic acid perfluorohexanoic acid perfluoroheptanoic acid perfluorooctanoic acid (also PFOA) perfluorononanoic acid perfluorodecanoic acid calcium chloride Cane Creek reservoir deionized water divinylbenzene equivalent background compound Environmental Protection Agency hydrochloric acid hour ideal adsorbed solution theory ionic strength internal standard liter liquid chromatagrpahy tandem mass spectrometry lowest observed adverse effect limit limit of quantitation milligram milliliter millimeter millimolar, mmol/l minute molar, mol/l xiv

18 NaCl NaHCO 3 NaOH ng NOAEL NOM NSF ORW OWASA sodium chloride sodium bicarbonate sodium hydroxide nanogram no observed adverse effect limit Natural organic matter National Sanitation Foundation Ohio River water Orange Water and Sewer Authority PAC powdered activated carbon PFC perfluorinated compound PFBS perfluorobutane sulfonate PFHS perfluorohexane sulfonate PFOA perfluorooctanoic acid (also C8) PFOS perfluorooctane sulfonate PHA Provisional Health Advisory pka acid-dissociation constant q SBA SO 4 2- S-PAC TOC UPW UV 254 WBA wk equilibrium solid-phase concentration strong base anion sulfate superfine powdered activated carbons total organic carbon ultrapure water ultraviolet absorbance at a wavelength of 254 nm weak base anion week µg microgram µl microliter xv

19 CHAPTER 1: INTRODUCTION AND OBJECTIVES MOTIVATION Perfluorinated compounds (PFCs) are organic compounds in which all carbon-hydrogen bonds are replaced with carbon-fluorine bonds. Two classes of PFCs that are of ecotoxicological and human health concern are perfluorocarboxylic acids and perfluorosulfonic acids. PFCs are surfactants used as active ingredients in stain repellents (e.g. Scotchgard) and firefighting foams. They are also used in the manufacture of non-stick coatings (e.g. Teflon) and water repellent fabrics (e.g. GoreTex). PFCs are released into the environment directly by fluorochemical production facilities and indirectly through consumer use of products. PFCs are of public health concern because of their toxicity and their persistence in the environment and the human body. Perfluorooctanoic acid (PFOA) and perfluorooctane sulfonate (PFOS) have been detected in the tissues of humans and wildlife even in remote regions of the globe. For example, measurable PFOS concentrations were found in the blackfooted albatross in the North Pacific Ocean, and ten-fold higher PFOS concentrations have been found in herrings from the highly industrialized North American Great Lakes region (Giesy and Kannan 2002). In addition, PFCs have been detected globally in human blood serum collected from both fluorochemical industry workers and the general population. Fluorochemical industry workers often have PFOS and/or PFOA blood levels that are an order of magnitude higher than those found in the general population (Lau et al 2007). Once 1

20 present in human whole blood or serum, PFOA and PFOS have half-lives of 3.8 and 5.4 years, respectively (Betts 2007). In terms of human health effects, a variety of adverse health conditions have been putatively associated with PFC exposure. Among them are decreased birth weight in newborns, increased risk of thyroid disease, chronic kidney disease, and decreased sperm count (Fei et al. 2007, Melzer et al. 2010, Shankar et al. 2011, Joensen et al. 2009). A recent study showed that PFC exposure in children from birth to age 5 years reduced the immune response to diphtheria and tetanus vaccinations to the point of antibody concentrations being below the level needed for long-term protection (Grandjean et al. 2012). Furthermore, some studies have shown a positive correlation between PFOA exposure and increased cholesterol, uric acid, and liver enzyme levels in adults; however, effects are not consistent across all studies (Steenland et al. 2010). In addition, a Johns Hopkins University study showed a correlation between PFOS and PFOA levels in newborn infants and a decreased birth weight and head circumference (Betts 2007). Overall, studies on health effects associated with human PFC exposure are still sparse and often inconclusive. As drinking water is one important exposure route, the EPA has issued a drinking water Provisional Health Advisory (PHA) for PFOA and PFOS at 0.4 and 0.2 μg/l, respectively (U.S. EPA 2009). The PHA was developed to protect consumers against adverse effects from short-term PFOA and PFOS exposure. In New Jersey, a health-based PFOA guideline of 0.04 μg/l was developed as a protective measure against lifetime PFOA exposure. PFOA levels of 0.04 μg/l are not uncommon in drinking water sources. In addition, the EPA is seeking 2

21 occurrence data for six PFCs [PFOA, PFOS, perfluoroheptanoic acid (C7), perfluorononanoic acid (C9), perfluorobutane sulfonate (PFBS), perfluorohexane sulfonate (PFHS)] in U.S. drinking waters as part of the proposed third Unregulated Contaminant Monitoring Rule (UCMR3). To date, little is known about the effectiveness of drinking water treatment processes for PFC removal. As more information about the effects of PFCs on human health becomes available, it is important to concurrently identify treatment methods that are capable of protecting human health by limiting PFC exposure through drinking water. RESEARCH OBJECTIVES The principal objective of this research was to assess the effectiveness of powdered activated carbon (PAC) adsorption and anion exchange processes for the removal of PFCs from drinking water sources. Apart from the more commonly studied PFOA and PFOS, the removal of eight additional PFCs [C4, perfluoropentanoic acid (C5), perfluorohexanoic acid (C6), C7, C9, C10, PFBS, PFHS] that are frequently detected in drinking water sources were studied at environmentally relevant concentrations. Specific objectives included to 1. Evaluate the effect of PAC base material and activation method on PFC removal. 2. Evaluate the effectiveness of superfine PAC (S-PAC) for PFC removal. S-PACs were produced from as-received PACs through a wet-milling process. Enhanced PFC adsorption kinetics were expected with S-PACs because of their smaller particle size. 3

22 3. Determine whether the blending of microporous and mesoporoous PACs was advantageous for PFC removal. Blending PACs may be advantageous because of the wide range of molecular weighs represented by the ten PFCs selected for this research. 4. Evaluate the effect of resin type on PFC removal by anion exchange. APPROACH Experiments evaluating PFC removal were conducted with five commercially available PACs prepared from different base materials (wood, coconut shell, lignite, bituminous coal) and with different activation methods (chemical, thermal). The five PACs were each ground into a superfine PAC (S-PAC) by a wet-milling process. PFC removal was also evaluated with four commercially available anion exchange resins. Resin types that were evaluated included strong and weak base anion resins containing quaternary amine and tertiary amine groups, respectively. Both acrylic and styrene-divinylbenzene copolymer resins were studied in gel and macroporous form. Experiments were conducted in three waters. The first was amended ultrapure water to evaluate PFC removal in the absence of natural organic matter (NOM). In ultrapure water, the effects of ph and ionic strength on PFC removal were evaluated. In addition, two drinking water sources were evaluated to determine background water matrix effects ( total organic carbon (TOC), ph, hardness, sulfate) on PFC removal: Cane Creek Reservoir water, which serves as a source for the Orange Water and Sewer 4

23 Authority (Carrboro, NC), and Ohio River water, which serves as a source for the Greater Cincinnati Water Works (Cincinnati, OH). 5

24 CHAPTER 2: BACKGROUND PERFLUORINATED COMPOUNDS Characteristics PFCs are organic compounds in which all carbon-hydrogen bonds are replaced with carbonfluorine bonds. Figure 2.1 shows the generic chemical structures of two important PFC classes, the perfluorocarboxylic acids and perfluorosulfonic acids. (a) (b) Figure 2.1 General chemical formulas for (a) perfluorocarboxylic acids (b) perfluorosulfonic acids. Figure 2.1 shows linear perfluorinated structures, but PFCs can also be branched (Lindstrom et al. 2011). PFCs are surfactants used as active ingredients in stain repellents (e.g. Scotchgard) and firefighting foams. They are also used in the manufacture of non-stick coatings (e.g. Teflon) and water repellent fabrics (e.g. GoreTex). PFCs are released into the environment directly by fluorochemical production facilities and indirectly through consumer use of products. PFCs are manufactured by electrochemical fluorination and telemorization. The production of PFCs by electrochemical fluorination produces a mixture of the target PFC 6

25 and homologues with additional or fewer perfluorinated carbons. PFCs work well as surfactants because of the hydrophilic acid head and hydrophobic carbon-fluorine tail that can not only repel water but fats and oils. The vapor pressures of PFOS and PFOA are 3.3x10-4 and 5 Pa, respectively (Rayne and Forrest 2010). The vapor pressures of PFCs decrease with increasing chain length, and low vapor pressure of PFCs limits volatization from aquatic environments. Physical/chemical properties of the ten PFCs studied in this project are summarized in Table

26 Table 2.1 Chemical properties of PFCs studied. PFC CAS # MW pka log D* (ph 7) log K OW Solubility at ph 7 (g/l) Intrinsic Solubility (g/l) Perfluorobutanoic acid (C4) a, 0.17 d a,-0.52 b a, 2.14 c 1000 a 1.1 a, 1.37 c Perfluoropentanoic acid (C5) a 0.64 a,0.09 b a, 2.81 c 1000 a 0.19 a, c Perfluorohexanoic acid (C6) a 1.24 a,0.7 b a, 3.48 c 260 a a, c Perfluoroheptanoic acid (C7) a 1.97 a,1.31 b a, 4.14 c 58 a a, c Perfluorooctanoic acid (C8) a 2.69 a,1.92 b a, 4.81 c 13 a, 3400 b 2.3 x 10-3a, 4.8x10-4c Perfluorononanoic acid (C9) a 3.42 a,2.57 b a, 5.48 c 3 a 5.1 x 10-4a, 6.2x10-5c Perfluorodecanoic acid (C10) a 4.15 a a, 6.15 c 0.67 a 1.2 x 10-4a, 8x10-6c Perfluorobutane sulfonate a a a, 1.82 c 999 a 0.9 a, c Perfluorohexane sulfonate a a a, 3.16 c 150 a a, c Perfluorooctane sulfonate a 1.01 a,2.45 b a, 4.49 c 7.5 a, 680 b 2.4 x 10-3a, 1x10-4c * ph-dependent octanol/water partition coefficient octano/water partition coefficient of molecular form a As listed in SciFinder Scholar (values were calculated with Advanced Chemistry Development (ACD/Labs) Software) b Rayne and Forrest 2010 c US EPA Estimation Programs Interface Suite for Microsoft Windows, v United States Environmental Protection Agency, Washington, DC, USA. d SPARC on-line calculator ( + Also denoted as PFOA 8

27 Large differences in the physical/chemical properties of PFCs with varying carbon chain length leave questions about how to predict the fate and transport of PFCs in the environment (Rayne and Forrest 2010). The log KOW ranges from for PFBS to for C10. Similarly log D at ph 7 ranges over about 6 orders of magnitude from for PFBS to 4.15 for C10. The aqueous solubility of the selected PFCs also varies widely. At ph 7, C4 and C5 are essentially miscible in water while the solubility of C10 is <1 g/l. Occurrence Studies have shown the presence of PFCs in the environment globally. PFCs have been detected in surface waters, groundwaters, wastewaters, and drinking water sources. Table 2.2 summarizes the results of a selection of these studies. 9

28 Table 2.2 Summary of studies detecting PFCs in water Location Type of PFOA PFOS Total PFC Source Water (ng/l) (ng/l) (ng/l) New Jersey Groundwater ND Post et al Surface water ND 39 Great Lakes Surface water Boulanger et al River Elbe basin, Treated WW ND Ahrens et al Germany Tennessee River, Surface water ND Hansen et al AL Cape Fear River basin, NC Surface water ND 287 ND up to 942 Nakayama et al Aurora, CO Surface water ND ND - Quinones and Snyder 2009 Pacific Treated WW Schultz et al Northwest Ruhr region, Germany Surface water ND ND up to 4385 Skutlarek et al Ruhr region, Drinking water ND 519 ND 22 up to 598 Skutlarek et al Germany Osaka, Japan Drinking water Takagi et al Fuxin, China Drinking water ND up to 2.7 Bao et al River Water up to 713 Bao et al Groundwater ND-0.73 up to 1400 Bao et al Decatur, AL Surface water ND-11,000 ND Lindstrom et al Levels of PFOA and PFOS range from non-detect to well above the U.S. Provisional Health Advisory. While the source of contamination in many areas is not clear, some of the studies were conducted near PFC production facilities (Decatur, AL; Fuxin, China). Table 2.2 also illustrates the degree of global contamination with PFCs detected in waters of the United States, Germany, Japan, China, and many other countries (Ahrens et al. 2009) Health Effects PFCs are of increasing public health concern because of their toxicity and persistence in the environment and human body. The persistence of a PFC in the environment and human body 10

29 is directly related to the length of the carbon chain. For example, perfluorobutanoic acid (PFBS) with 4 carbons can be eliminated from the human body in 4 months, while PFOA and PFOS require 3.8 and 5.4 years, respectively, for elimination (Betts 2007). In the United States, PFOS and PFOA concentrations measured in whole blood samples from the general population were as high as 73.2 ng/ml and 1656 ng/ml, respectively. Globally, PFOS has been detected in whole blood at concentrations as high as 116 ng/ml in Poland (Lau et al. 2007). Studies have shown that the blood serum concentration in humans in the industrialized world ranges from 2-8 ng/ml PFOA. A study conducted with California women monitored PFC levels from Samples were taken from different groups of US women in 1960, 1980, and 2010 and compared to National Health and Nutrition Examination Survey results (Wang et al. 2011). PFOS showed a decrease over the fifty years, from 45.9 ng/ml to 9.44 ng/ml. The decrease from is consistent with the phase out of PFOS manufacturing in For PFOA, there was an increase from 0.3 ng/ml to 3.17 ng/ml from , but a decrease to 2.21 ng/ml in The steady decrease is PFOS levels in human blood accompanied by only a minimal decrease in PFOA levels may indicate sources other than fluorochemical plants. In addition, a steady increase in longer chain PFCs like perfluorononanoic acid and perfluorodecanoic acid occurred over the fifty years. The consistent build-up of longer chain PFCs could be because of their continued production and emission and the longer half-lives compared to PFOA. Unlike other persistent organic pollutants like polychlorinated biphenols which are lipophilic and accumulate in fatty tissue, PFCs accumulate in organs like liver, kidney, muscle, and attach to the serum proteins in blood. While it has been established that PFCs are present in the environment and humans, 11

30 to date, epidemiological studies on human health effects of PFCs are often inconclusive. Research efforts have provided information of the toxicokinetic effects in animals such as rats and monkeys, but these differ significantly between species. For example the half-life of PFOA is 3.8 years in humans, but ranges from 4 hours in female rats to 6 days in male rats to days in mice (Lindstrom et al. 2011). The difference in toxicokinetics makes extrapolation of animal PFC health effect research to humans difficult. Presented here are human and animal epidemiological studies that point to possible adverse health effects of PFC exposure. Studies have shown elevated exposure to PFCs, measured by a comparison of PFCs levels in blood serum, in children from birth to age 5 years reduced the immune response to immunizations for diphtheria and tetanus to the point of antibody concentrations being below the level needed for long-term protection. A two-fold increase in PFOS exposure was associated with a decrease in antibody concentration in blood serum by 39% at age 5 years. A cumulative two-fold PFC exposure in children was associated with a decrease of 49% of antibody concentrations in blood serum, with PFOS and PFOA representing the main culprits (Grandjean et al. 2012). The same study also showed that postnatal PFC exposure from birth until age 7 years was more strongly associated with decrease in antibody concentration than prenatal exposure. A study of Danish pregnant women found a negative correlation between birth weight and maternal plasma PFOA concentrations (Fei et al. 2007). In addition, a Johns Hopkins study 12

31 showed a correlation between higher PFOS and PFOA levels in newborns and lower birth weight and head circumference (Betts 2007). Recent studies also showed a positive correlation between PFOA exposure and increases in cholesterol, uric acid, and liver enzymes in adults (Steenland et al. 2010). A study on data collected as part of the National Health and Nutrition Survey from showed a correlation between adults with thyroid disease and PFOA and PFOS concentrations in blood serum (Melzer et al. 2010). Women with PFOA blood serum concentrations 5.7 ng/ml were more than twice as likely to have thyroid disease as women with 4 ng/ml. A similar increase in thyroid disease was observed in men with 7.3 ng/ml PFOA in their blood serum compared to those with 5.2 ng/ml. For PFOS, blood serum concentrations 36.8 ng/ml compared to 25.5 ng/ml resulted in men being over 2.5 times more likely to develop thyroid disease. No correlation between PFOS concentration and thyroid disease was observed in women (Melzer et al. 2010). A study of U.S. adults showed a positive association between increased PFOA and PFOS levels and chronic kidney disease. Adults with >5.9 and >29.5 ng/ml blood serum concentrations of PFOA and PFOS, respectively, were almost two times as likely to have chronic kidney disease (Shankar et al. 2011). Increased levels of PFCs have also been linked to insulin resistance and high serum uric acid levels which are also associated with the development of chronic kidney disease. A study of sperm count in Danish men found that elevated levels of PFOA and PFOS in blood serum resulted in significantly lower spermatozoa levels in semen. Average PFOS and 13

32 PFOA blood serum concentrations were 4.9 and 24.5 ng/ml, respectively. Men in the highest quartile based on PFOA and PFOS concentration in blood serum had 6.2 million total morphologically normal semen compared to 15.5 million in the lowest PFOA and PFOS concentration group (Joensen et al. 2009). Studies have also been conducted to determine the potential adverse health effects of PFC exposure in animals. Exposure to PFOA in female rats resulted in decreased body weight gain, decreased erythrocytes, and decreased hemoglobin concentration (Post 2007). A study of PFOA exposure in Rhesus monkeys caused an increase in liver weight. A study of rats exposed to PFOA showed an increased incidence of Leydig cell adenomas in male rats and an increase in mammary fibro adenomas in female rats. Male rats exposed to PFOA also had an increased incidence of hepatic and pancreatic adenomas and carcinomas (Post 2007). Studies of mice showed prenatal exposure to PFOS and PFOA caused increased incidence of neonatal mortality. Mice were born appearing to be healthy, but would die shortly after birth (Betts 2007). Studies in animals have shown that PFOA and PFOS can bind to the peroxisome proliferator-activated receptor (PPAR), which is receptor associated with carcinogenesis. The activation of PPAR is believed to be related to the cause of liver and pancreatic tumors in mice and the effects seen on fetal growth and immune response (Betts 2007). While there is a plethora of studies reporting adverse health effects of PFCs in animals, extrapolating these results to humans is controversial because of the considerably shorter PFC half-life in animals compared to humans. 14

33 Health Advisories As drinking water is one important exposure route, the EPA has issued a drinking water Provisional Health Advisory (PHA) for PFOA and PFOS at 0.4 and 0.2 μg/l, respectively (US EPA 2009). The PHA was developed to protect consumers against adverse effects from short-term PFOA and PFOS exposure. In New Jersey, a health-based PFOA guideline of 0.04 μg/l is in effect to protect against adverse health effects from a lifetime of exposure to drinking water containing PFOA (Post et al. 2009). To derive the 0.04 μg/l PFOA guideline, New Jersey used a method relating the external PFOA dose (drinking water concentration) to the blood level of PFOA in humans. A concentration factor of 100 from drinking water to blood was used. By evaluating toxicity studies in a variety of animal types (adult rats, pregnant female rats, non-human primates, rat pups) that modeled animal blood levels after long-term exposure to PFOA, no observed adverse effect levels (NOAEL) and lowest observed adverse effect levels (LOAEL) were correlated to blood PFOA concentration in animals. Different toxicity endpoints were used in each study, such as decreased body weight and hematology in adult female rats and increased liver weight and possible mortality in adult non-human primates. After applying uncertainty factors that were study-specific, a target human blood level was extrapolated from the animal blood level. To this target human blood level, the concentration factor of 100 was applied. In addition, it was assumed that the drinking water contribution to PFOA exposure was 20%. The six health-based concentrations calculated ranged from μg/l, and the most conservative of these values was selected for the health-based standard (Post 2007, Post et al. 2009). PFOA levels of 0.04 μg/l are not uncommon in drinking water sources (Table 2.2). 15

34 Effectiveness of Conventional Water Treatment Processes for PFC Removal The strength of the carbon-fluorine bond makes PFCs resistant to conventional water treatment methods. Multiple studies showed that coagulation, flocculation, and filtration are ineffective for PFC removal (e.g, Vecitis et al. 2009, Quinones and Snyder 2009). Oxidation processes such as chlorination, ozonation, and advanced oxidation are also ineffective. Fluorine is the most electronegative element, making it thermodynamically unfavorable for any oxidant to break the carbon-fluorine bond and release a fluoride ion (Vecitis et al. 2009). Tagaki et al. (2008) presented results of a sampling survey of PFCs in 14 Japanese water treatment plants. Only a treatment plant utilizing activated carbon that was changed regularly was effective at removing PFCs. Plants utilizing chlorination, ozonation, slow-sand filtration, and rapid filtration were not effective at removing PFOS and PFOA. Reverse osmosis and nanofiltration have been shown to be effective. Reverse osmosis resulted in >99% rejection of PFOS and nanofiltration resulted in 90-99% removal (Tang et al. 2007). The effectiveness of reverse osmosis treatment was also observed by Quinones and Snyder (2009), but high energy requirements make reverse osmosis a costly option for PFC removal. ACTIVATED CARBON ADSORPTION Powdered Activated Carbon Activated carbons are classified as PAC by AWWA Standard B if not less than 90% by mass passes through a 44-µm sieve. Wood-based carbons are an exception and are classified as PAC if not less than 60% by mass passes through a 44-µm (Summers, Knappe, 16

35 and Snoeyink 2011). PAC can be produced from a variety of organic feedstocks, such as wood, coconut shells, bituminous coal, and lignite. The raw material is turned into a char by pyrolytic carbonization and then oxidized to develop the internal pore structure. This internal pore structure is what provides the large surface area that makes activated carbon effective for water treatment. Activation, the development of the internal pore structure, is commonly accomplished in two ways: chemically or thermally. Thermal activation occurs at temperatures between 800 and 900 C with oxidizing gases such as steam and/or carbon dioxide. Chemical activation is accomplished by heating the raw material with phosphoric acid in the absence of oxygen. Depending on the raw material and activation method, PACs vary in surface chemistry, surface area, and pore size distribution. The surface chemistry of activated carbons is defined by the presence of elements other than carbon. Carbon typically has an elemental composition of ~90% carbon, 6-7% oxygen, 1 % sulphur, 0.5% nitrogen, and 0.5 % hydrogen, with the remainder being ash. Oxygen is the most important non-carbon element due to its effect on hydrophilicity and surface charge. Oxygen will occur in the form of carboxylic acid groups, phenolic hydroxyl groups, and quinone carbonyl groups. The surface area is commonly measured by the Brunauer-Emmett-Teller (BET) nitrogen adsorption method. Typical BET surface areas for PAC are m 2 /g. Pore size distribution is divided into three classifications: micropores, mesopores, and macropores. Micropores have pore diameters less than 2 nm, mesopores between 2 and 50 nm, and macropores have diameters greater than 50 nm. Pore size distribution will affect the type of compound that can be adsorbed and the kinetics of adsorption (Summers, Knappe, and Snoeyink 2011). 17

36 Effectiveness of Activated Carbon Adsorption for PFC Removal Bench scale studies have shown activated carbon adsorption to be effective for PFOS and PFOA removal. Column tests with granular activated carbon (GAC) showed GAC was effective at removing >90% of PFOS from a 5 mg/l PFOS solution in deionized water for 25,000 bed volumes (Senevirathna et al. 2010). Hansen et al. (2010) confirmed the effectiveness of activated carbon for the removal of nine PFCs from highly contaminated well water, in which PFOA and PFOS were present at concentrations of ~1,400 µg/l. Hansen et al. (2010) observed that PFC adsorbability increased with increasing carbon chain length, and that a sulfonic acid of a given chain length was more adsorbable than the corresponding carboxylic acid. Greater than 95% removal of PFOS and PFBS was achieved with ~1,700 mg/l GAC in 15 hours from solutions with initial PFOS and PFBS concentrations of 1 mg/l (Carter and Farrell 2010). ION EXCHANGE Ion exchange is a process in which exchangeable ions on a resin are displaced by ions in solution. Ion exchange is typically used in drinking water treatment for water softening as well as for the removal of sulfate, nitrate, bromide, and NOM. Anion Exchange Resins The resin is an insoluble support matrix for the exchangeable ions and can be natural or synthetic. Synthetic resins tend to be more durable and can be regenerated with a salt or acid/base solution. Synthetic resins are in the form of spherical beads and come in a variety 18

37 of bead sizes.typical sizes are 16x40, 16x50, and 20x50 U.S. mesh. The backbone of the resin is made up of a network of cross-linked polymers that are covalently bonded through fixed functional groups. Cross-linking is accomplished with the use of divinylbenzene. The most common polymers for making the resin matrix are polyacryl and polystyrene. Figure 2.2 shows the generic structure of an anion exchange resin with crosslinked polymer backbone. Figure 2.2 Strong-base anion exchange resin (Source: Clifford, Sorg, and Ghurye, 2011) Resins are classified as gel type or macroporous resins. Gel-type resins typically have approximately 8% cross-linking, which results in an open pore structure that can result in faster kinetics of ion exchange. Macroporous resins typically have 20-25% crosslinking. The degree of cross-linking affects the resistance of the resin to swelling or shrinking. Macroporous resins retain their pore structure when dried, whereas gel-type resins lose their 19

38 pore structure upon drying. Gel-type resins typically have very low BET surface areas while macroporous resins have BET surface areas of m 2 /g. Macroporous resins are more resistant to swelling, and the moisture content of these resins is lower. Gel-type resins have a higher water content, and the more open pore structure allows for more swelling. Swelling due to ion exchange can be reversible, but repeated swelling and shrinking can cause beads to fracture. Resin quantities are often measured on a volumetric basis, which includes water within the resin pores. For gel-type resins with their more open pore structure and higher water content, the volume of actual resin measured by volume will be less than with a macroporous resins. Resin capacities are expressed in equivalents per liter, which is dependent on the quantity of functional groups per wet-volume of resin. The wet-volume depends on the moisture content of the resin (MWH, 2005). Anion exchange resins are grouped into strong base anion (SBA) and weak base anion (WBA) resins. The distinction is in the pk values of the functional groups, which are >13 for strong base and for weak base anion exchange resins (MWH, 2005). A common functional group in SBA resins is a quaternary amine group, which carries a positive charge over almost the entire ph range of 0 to 14. In WBA resins, tertiary amines are the most common functional group, which does not become appreciably protonated until the ph falls below the pk. The general exchange reactions for a SBA and WBA resins are presented in Equations 2.1 and 2.2, respectively. [ ( ) ] [ ( ) ] Equation

39 [ ] [ ] Equation 2.2 In equation 2.1, [R 4 (CH 3 )N + ] represents the anion exchange site of strong base anion exchange resin which is originally in the chloride form. For strong base resins, the anion in solution (e.g. PFC - ) will exchange with a chloride ion on the functional group. In equation 2.2, [R 3 N] represents the tertiary amine functional group of the WBA resin. It is only ionized in the acidic ph region, at which point it is able to serve as an anion exchange site (Clifford, Sorg, and Ghurye, 2011). Effectiveness of Anion Exchange for PFC Removal Ion exchange resins have shown promise for PFC removal. Macroporous styrenedivinylbenzene (DVB) resins were effective for PFOS removal in column tests (Senevirantha et al. 2010). Polyacrylic quaternary amine SBA resin at a dose of ~1.7 g/l achieved >95% removal of PFOS and PFBS in four hours (Carter and Farrell 2010). Similar removal trends were observed with ion exchange resins as with activated carbon. PFOS uptake on resins was greater than PFBS, perhaps due to the greater hydrophobicity of the longer fluorinated PFOS chain (Carter and Farrell 2010). Resin type has also been shown to play a role in PFOS removal. Polyacrylic resins exhibited faster uptake kinetics and higher capacity compared to polystyrene resins, regardless of whether the resins were tested in gel-type or macroporous form (Deng et al. 2010). The faster uptake kinetics and higher capacity were attributed to the more hydrophilic properties of the polyacrylic resin. The hydrophilic properties of the resin aid in the movement of water into the pores, which aids in the transport of PFCs into the pores. Differences in gel-type and macroporous resins have been observed in polystyrene- 21

40 DVB resins. Faster uptake kinetics were observed with macroporous polystyrene-dvb resins compared to gel-type polystyrene-dvb resins (Deng et al 2010). The PFC removal effectiveness of ion exchange resins can be affected by the presence of other ions in solution. Deng et al. (2010) showed the presence of sulfate decreased the availability of ion exchange sites for PFOS. SUMMARY Based on the literature reviewed here, it is apparent that common water treatment processes such as coagulation, sedimentation, and filtration are not effective for PFC removal. Among the processes that show promise for PFC removal (activated carbon adsorption, anion exchange, nanofiltration, reverse osmosis), PAC adsorption and anion exchange were selected for a more detailed evaluation. The research objectives of this study were designed to fill knowledge gaps by studying (1) ten perfluorinated compounds (rather than the commonly studied PFOA and PFOS), (2) drinking water relevant concentrations (500 ng/l for each PFC rather than the higher concentrations used in prior studies), and (3) superfine PACs were for their effectiveness to remove PFCs. 22

41 CHAPTER 3: MATERIALS AND METHODS MATERIALS Water Experiments were conducted in amended ultrapure water (UPW), Cane Creek Reservoir water (Carrboro, NC), and Ohio River water (Cincinnati, OH). Cane Creek Reservoir (CCR) water serves as a drinking water source for the Orange Water and Sewer Authority (OWASA) and Ohio River water (ORW) for the Greater Cincinnati Water Works. For batch kinetic tests, the ionic strength of UPW was adjusted with NaHCO 3 and NaCl to 0.02, 0.01, and and adjusted with HCl or NaOH to ph 5.5, 7.0, and 8.5. For isotherm tests, UPW was amended with 15 mm NaHCO 3 and 5 mm NaCl (ionic strength = 0.02) and adjusted with HCl to ph 7.0. Table 3.1 summarizes the water quality parameters for the natural waters. Two drums of CCR were used for experiments. Unless otherwise noted, CCR data presented throughout this thesis resulted from experiments conducted with CCR water in drum 1. 23

42 Table 3.1 Water quality parameters for drinking water sources Cane Creek Reservoir (CCR) Ohio River Drum 1 Drum 2 Water (ORW) ph TOC (mg/l) UV Alkalinity (mg/l as CaCO 3 ) Hardness (mg/l as CaCO 3 ) 35 NM 127 Turbidity (NTU) Sulfate (mg/l) NM: not measured Activated carbons Experiments were conducted with five commercially available PACs in as-received form. PACs were given the labels A-E as shown in Table 3.2. Four PACs were thermally activated and prepared from difference base material. Two wood-based PACs were prepared by two activation methods, chemical and thermal. PACs were dried at 105 C for 24 hours and then stored in a desiccator until use. Table 3.2 As-received PACs PAC Base Material Activation A coconut shell thermal B lignite thermal C wood thermal D wood chemical E bituminous coal thermal * based on information provided by manufacturer In addition, experiments were conducted with five superfine PACs (S-PACs.S-PACs were obtained by wet milling as-received PACs to a target d 50 of ~1 µm (NETZSCH Premier Laboratory). For S-PACs, the nomenclature shown in Table 3.3 was used. 24

43 Table 3.3 S-PAC Inventory Starting PAC A B C D E S-PAC Nomenclature S-A S-B S-C S-D S-E After wet-milling, S-PACs were received in slurry form with ~7-11% solids. The concentrated S-PAC slurry was diluted approximately 1:20 by volume and the S-PAC concentration measured by drying three 10-mL aliquots at 105 C for hours and weighing the mass of solids remaining. Diluted S-PAC slurries ranged in concentration from 2 20 mg/ml. The dilute S-PAC slurry was used for dosing kinetic tests, isotherm, and jar tests. Carbon characteristics such as surface area, pore volume distribution, and elemental analysis are presented in Chapter 4. Anion Exchange Resins Experiments were conducted with four commercially available anion exchange resins that are NSF-certified for use in drinking water treatment. Resins were given the labels 1-4 as shown along with their properties in Table 3.4. All strong base anion (SBA) resins were in the chloride form and the weak base anion (WBA) resin was in the free base form. Resins were prepared by rinsing resin beads with at least 20 bed volumes of deionized water through a column to remove fines. Rinsed resins were stored at room temperature in deionized water until use. 25

44 Table 3.4 Anion exchange resin properties Resin Matrix Functional Group Total Capacity 1 Polyacrylic Macroporous Quaternary amine (strong base) 2 Styrene- Gel Quaternary amine Divinylbenzene (strong base) 3 Styrene- Macroporous Quaternary amine Divinylbenzene (strong base) 4 Acrylic Gel Tertiary amine Divinylbenzene (weak base) Resin Mass/Volume (mg/ml) 0.52 eq/l eq/l eq/l eq/l 275 Perfluorinated Compounds The ten PFCs studied in this project are listed in Table 3.5. The PFCs were chosen based on occurrence and the availability of an LC-MS/MS method. The pk a values of the PFCs listed in Table 3.5 range from to 0.52, meaning that the PFCs were present in their anionic carboxylate and sulfonate forms over the ph range evaluated in this study (5.5 to 8.5). Throughout this thesis the carboxylic acids will be referred to as C4-C10 and the sulfonates will be referred to by the abbreviations PFBS, PFHS, and PFOS. Table 3.5 Perfluorinated compounds selected for this study Compound Formula CAS Number Molecular Weight Perfluorobutanoic acid (C4) C 4 HF 7 O Perfluoropentanoic acid (C5) C 5 HF 9 O Perfluorohexanoic acid (C6) C 6 HF 11 O Perfluoroheptanoic acid (C7) C 7 HF 13 O Perfluorooctanoic acid (C8, PFOA) C 8 HF 15 O Perfluorononanoic acid (C9) C 9 HF 17 O Perfluoropentanoic acid (C10) C 10 HF 19 O Perfluorobutane sulfonate (PFBS) C 4 HF 9 SO Perfluorohexane sulfonate (PFHS) C 6 HF 13 SO Perfluorooctane sulfonate (PFOS) C 8 HF 17 SO

45 A stock standard solution containing 5 µg/ml each of the ten PFCs in methanol was purchased (Wellington Labs, Ontario, Canada). Stock solutions for calibration curves and spiking of reactor contents were prepared from neat PFC standards (Wellington Labs). Neat standards were weighed to produce individual PFC stocks of 10,000 ng/µl in methanol. A combined PFC stock of 1,000 ng/ul for each PFC was prepared by combining 1 ml of each individual 10,000 ng/µl PFC stock. Dilutions of the combined stock produced calibration stocks with individual PFC concentrations of 0.01 ng/µl and 0.1 ng/µl in methanol and a reactor dosing stock containing 10 ng/µl in methanol. The stocks made from neat compounds were compared to the purchased PFC mixture to confirm concentrations. Six internal standards [mass labeled C4 (MC4), mass labeled C6 (MC6), mass labeled C8 (MC8), mass labeled C10 (MC10), mass labeled PFHS (MPFHxS), and mass labeled PFOS (MPFOS)] were purchased from Wellington Labs each at a concentration of 50 µg/ml. A 20 µl aliquot of each internal standard was added to 10 ml of methanol to produce an internal standard stock of 0.1 ng/µl. Each standard and sample was spiked with 25 µl of the internal standard mixture. METHODS (S-)PAC Batch kinetic tests Batch kinetic tests were conducted to evaluate PFC removal by (S-)PACs from UPW, CCR, and ORW. Batch kinetic tests were conducted in amber glass bottles containing 0.45 L of non-filtered water, unless otherwise noted. Waters were spiked with 23 μl of a stock solution containing 10 ng/μl of each PFC in methanol. The targeted initial concentration was

46 ng/l for each PFC. (S-)PAC was added in slurry-form and mixed with a PTFE-coated magnetic stir bar. PTFE-coated magnetic stir bars were rinsed in methanol prior to each use to avoid potential PFC contributions. A blank reactor was stirred with a PTFE-coated stir bar and contents analyzed for PFCs to confirm the stir bar did not contribute PFCs to the reactor contents. After (S-)PAC addition, samples were taken at 5, 15, 30, 60, and 120 minutes and analyzed for PFC concentrations. An initial sample was taken prior to (S-)PAC addition (time 0 sample). For batch kinetic tests conducted in natural waters, additional samples were taken at time 0 and 120 minutes and analyzed for UV 254 absorbance and ph. Prior to analysis, each sample was filtered through a 0.45-µm glass microfiber membrane filter (25 mm GD/X, Whatman) to remove (S-)PAC and other particulate matter (S-)PAC Isotherm tests Isotherm tests were conducted to determine the equilibrium uptake capacity of (S-)PACs for each of the ten PFCs. Experiments were completed with carbons C and S-C in UPW, CCR, and ORW. Amber glass bottles containing 0.45 L of non-filtered water were spiked with the PFC stock solution to obtain initial concentration targets of 500 ng/l for each PFC. In UPW, (S-)PAC doses ranged from 0.3 to 500 mg/l, in CCR from 7 to 500 mg/l, and in ORW, from 4 to 750 mg/l were evaluated. At least half of the (S-)PAC doses were tested in duplicate. Bottles were mixed on a rotary tumbler and sampled after contact times of 2 and 3 weeks. CCR and ORW isotherm samples were analyzed for PFC concentrations, UV 254 absorbance, and ph. Only PFC concentrations were measured for experiments conducted in UPW. 28

47 An initial concentration sample was taken from each bottle prior to (S-)PAC addition (time 0 sample). Bottles containing PFCs but no carbon served as adsorbent-free blanks. A comparison of initial PFC concentrations and PFC concentrations measured after 2 and 3 weeks of tumbling showed that no measurable PFC losses to bottle surfaces occurred during equilibration (Appendix A). Isotherm Modeling Approach The Freundlich model, Equation 3.1, was used to describe the isotherm data obtained for each PFC. q = K*C 1/n Equation 3.1 To determine K and 1/n, the equilibrium solid-phase concentration (q, ng/g) was plotted as a function of equilibrium aqueous-phase concentration (C, ng/l) on a log-log plot. In addition, the simplified ideal adsorbed solution theory (IAST) combined with the equivalent background compound (EBC) method was used to describe PFC adsorption isotherm data in CCR and ORW. The EBC approach assumes that only a fraction of the NOM, the EBC, competes for adsorption sites, and that the adsorption isotherm of this fraction can also be described by Freundlich isotherm parameters. Using this approach, the difference in the PFC adsorption capacity of the (S-)PAC in UPW and in natural water can be modeled (Qi et al. 2007). For a trace compound in an ideal batch reactor, the relationship between carbon dose and the fraction of compound remaining is presented in equation

48 [ ] [ ] Equation 3.2 In equation 3.2, C C is the carbon dose in units of mg/l. K 1 ((ng/mg)(l/ng) 1/n1 )and n 1 are Freundlich parameters for the trace compound in UPW. C 2,0 (µm) is the initial concentration of the EBC. The parameter n 2 is the inverse of the Freundlich exponent for the EBC. C 2,0 and n 2 are not known. To determine C 2,0, the percent difference between the predicted fraction of PFC remaining and what was achieved experimentally in natural water was minimized while varying C 2,0 using Excel Solver. For the calculated model fraction, n 2 was assumed to be equal to n 1 of the PFC is UPW. C 2,0 and n 2 are not independent. An alternative approach would be to combine the expression (K 1 n 1 /n 2 )C ((1-n1)/n1) 2,0 into one parameter A (Qi et al. 2007). Equation 3.2 can be rearranged to predict the (S-)PAC dose required to achieve a desired PFC removal percentage (Equation 3.3). [ [ ] ( ) ( ) ] Equation 3.3 (S-)PAC Jar Tests Jar tests were conducted to evaluate the removal of PFCs with (S-)PACs and alum coagulation in CCR water. Experiments were conducted in a programmable jar testers (Phipps and Bird Model PB-900, Virginia) with 2 L square acrylic jars (B-KER, Phipps and Bird). Jars were filled with 1.89 L of CCR water. The addition of 110 ml of a 1.11 g/l alum 30

49 stock solution brought the total reactor volume to 2 L and the alum concentration to 55 mg/l. This alum dose is typically used by OWASA for treatment of CCR water. All jar tests were conducted at ph 6.2 (typical OWASA coagulation ph) unless otherwise noted. NaOH was added prior to alum addition in order to achieve the target ph of 6.2 once alum was added. Fine tuning of ph with NaOH and HCl was conducted immediately after alum addition. Waters were spiked to a targeted concentration of 500 ng/l for each PFC by dosing 100 µl of a 10 ng/µl PFC stock. (S-)PACs were added in slurry form. A jar containing water dosed with PFCs but no carbon was also stirred and analyzed for PFC concentrations to confirm that PFC loss to the acrylic walls were negligible for all PFCs except PFOS and C10, for which losses were 16 and 29% respectively (Figure A.2). Jar tests were conducted to evaluate the PFC removal obtained by different (S-)PAC types, (S-)PAC doses, coagulation ph values, and points of (S-)PAC addition relative to that of alum. (S-)PACs B, S-B, C, and S-C were evaluated in jar tests, and tested (S-)PAC doses were 15, 25, and 50 mg/l for (S-)C and 15 and 25 mg/l for (S-)B. Coagulation ph effects were evaluated with (S-)C at ph 5.5 and 7.0. The timing of (S-)PAC addition was evaluated by dosing (S-)C (a) 5 minutes prior to the addition of alum, (b) at the same time as alum, and (c) 9 minutes into flocculation. Jars containing alum only and (S-)PAC only were also evaluated. Scenario (a) simulates the addition of (S-)PAC at the intake or to a dedicated (S-)PAC mixing tank, and scenario (b) simulates the most common (S-)PAC addition point. Scenario (c) was evaluated to assess whether (S-)PAC addition to water in which a substantial fraction of the raw water turbidity and NOM is captured in the Al(OH) 3 floc, 31

50 which may lead to improved (S-)PAC performance. Table 3.6 shows the mixing regime used for all jar tests. Table 3.6 Jar test mixing regime Mixing Speed Time (rpm) Rapid Mix seconds, 330 seconds* Flocculation minutes Sedimentation 0 10 minutes *for test in which (S-)PAC was added 5 min prior to alum PFC and UV 254 samples were collected prior to (S-)PAC and alum addition and after 10 minutes of sedimentation. Turbidity samples were collected prior to (S-)PAC and alum addition and after 3.5 and 10 minutes of sedimentation. Prior to PFC and UV 254 analysis, samples were filtered through a 0.45-µm glass microfiber membrane filter to remove (S-)PAC and other particulate matter. Batch Kinetic Tests Conducted with Anion Exchange Resin Batch kinetic tests were conducted to evaluate PFC removal from UPW, CCR, and ORW with four anion exchange resins. Kinetic tests were conducted in 2 L acrylic jars using a jar tester. Jars were filled with 1 L of water, and reactor contents were spiked with 50 µl of a 10 ng/µl PFC stock solution to reach the targeted concentration of 500 ng/l for each PFC. After spiking of PFCs, jars were mixed at 100 rpm with the jar tester paddle. Resins were dosed at concentrations of 1, 5 and 10 ml/l. Resin doses were measured in 10-mL graduated cylinders. After resin addition, samples were taken at 5, 15, 30, 60, and 120 minutes and analyzed for PFC concentrations. An initial sample was taken prior to resin addition (time 0 sample). For batch kinetic tests conducted in natural waters, additional samples were taken at 32

51 time 0 and 120 minutes and analyzed for UV 254 and ph. Prior to analysis, each sample was filtered through a 0.45-µm glass microfiber membrane filter to remove resin and other particulate matter PFC analysis PFC concentrations were measured by liquid chromatography tandem mass spectrometry (LC-MS/MS) at the National Exposure Research Laboratory of the Environmental Protection Agency (EPA) in Research Triangle Park, NC. An Agilent 1100 Series LC pump and PE Sciex API 3000 LC/MS/MS system equipped with a FlouroFlash HPLC column (4.6 mm i.d. x 50 mm, Flourous Technologies Inc.) was used for PFC separation. The gradient method shown in Table 3.7 was used to elute PFCs from the HPLC column. Mobile phase A was 2 mm ammonium acetate in DI water with 5% methanol. Mobile Phase B was 2 mm ammonium acetate in methanol with 5% DI water. Time (min) Table 3.7 LC Gradient Method for PFC Analysis Mobile Phase A % (v/v) Mobile Phase B % (v/v) Flow Rate (ml/min) Mass spectrometer transitions for PFC analytes and internal standards used are shown in Table

52 Analyte or Internal Standard Table 3.8 Mass transitions for PFCs and internal standards MS/MS Transition Declustering Potential (V) Focusing Potention (V) Collision Energy Collision cell exiting potential C > C > C > C > C > C > C > PFBS > PFHS > PFOS > Perfluoro-n-[1,2,3,4-13C4 ]butanoic acid (MC4) Perfluoro-n-[1,2-13 C 2 ]hexanoic acid (MC6) Perfluoro-n-[1,2-13 C 2 ]octanoic acid (MC8) Perfluoro-n-[1,2-13 C 2 ]decanoic acid (MC10) Sodium perfluoro-1-hexane[ 18 O 2 ]sulfonate (MPFHxS) Sodium perfluoro-1-[1,2,3,4-13c4]octane sulfonate (MPFOS) > > > > > >

53 Calibration curves were produced by analyzing standards with concentrations of 10, 25, 50, 100, 250, 500, and 750 ng/l. Calibration standards were prepared in the same water that was used in a given experiment (buffered UPW, CCR, or ORW). The limit of quantitation (LOQ) for all compounds was 10 ng/l for all PFCs except C10, for which the LOQ was 25 ng/l. Calibration curves were produced with seven standards ranging from 10 to 750 ng/l. Water was spiked with calibration stocks containing either 0.1 ng/µl or 0.01 ng/µl of each PFC, depending on the targeted standard concentration, and filtered with a 0.45-μm glass fiber syringe filter. Calibration standards were analyzed at the beginning and end of LC runs and averaged to account for instrument drift. All samples and calibration standards were filtered through a 0.45-μm glass fiber syringe filter into a 15-mL Becton-Dickinson polypropylene vial. After filtration, samples and standards were spiked with 25 µl of the internal standard mixture and 4 µl formic acid, which was used to lower the ph to below 2. A blank with no PFCs added was also prepared with the calibration standards and plotted as a calibration curve point. Mass spectrometer results were quantitated based on analyte peak area counts. Internal standard area count was used to verify an error did not occur during sample injection if the results were unusual. Chromatograms for each of the ten PFCs and six internal standards are presented in Appendix B. Appendix C presents representative calibration curves for each PFC. 35

54 Total Organic Carbon Analysis The total organic carbon (TOC) concentration was measured by high-temperature combustion (Shimadzu TOC-VCSN) according to Standard Method 5310B. The UVabsorbing organic constituent concentration was measured at a wavelength of 254 nm, according to Standard Method

55 CHAPTER 4: RESULTS AND DISCUSSION PFC REMOVAL BY ACTIVATED CARBON (S-)PAC Physical and Chemical Characteristics Surface area, micropore and mesopore volumes, and elemental composition were determined for each of the ten (S-)PACs. Physical/chemical properties are presented in Tables 4.1 and 4.2 (Dunn, 2011). Table 4.1 shows that BET surface areas and micropore volumes were similar for a given PAC/S-PAC pair. In contract, mesopore volumes increased after wet-milling for all carbons except (S-)D. An increase in the volume of mesopores, the pores responsible for transport of the adsorbate into the carbon particle, suggest wet-milling as-received carbons may increase the kinetics of adsorption beyond what can be expected based on size reduction alone. Wet-milling as-received PACs resulted in a small decrease in carbon content, while the hydrogen and nitrogen contents remained similar (Table 4.2). In contrast, the oxygen content increased by 37-66% upon wet-milling of as-received PACs. The ph PZC results illustrate that the as-received forms of carbons A-C were basic (ph PZC ranged from ) while carbons D and E were acidic. The lowest ph PZC of 4.9 was obtained for the chemically activated wood-based PAC (carbon D). Wet-milling of all five as-received PACs moved the ph PZC closer to neutral. 37

56 Table 4.1 Surface area and pore volume analysis (S-) PAC Base Material BET Surface Micropore Mesopore Mean diameter Diameter Area (m 2 /g) volume (cc/g) volume (cc/g) (µm) Range (µm) A Coconut shell S-A B Lignite coal S-B C Wood S-C D Wood S-D E Bituminous coal S-E

57 Table 4.2 Elemental analysis (S-) PAC Base Material Ash Content (%) Carbon (%) Hydrogen (%) Nitrogen (%) Oxygen (%) Sum (%) A Coconut shell S-A B Lignite coal S-B C Wood S-C D Wood S-D E Bituminous S-E coal ph PZC 39

58 (S-)PAC Batch Kinetic Tests The objective of the batch kinetic tests was to compare PFC adsorbability and adsorption kinetics on (S-)PACs with different physiochemical characteristics. In addition, batch kinetics tests were conducted to evaluate the effect of background water constituents on PFC removal. Adsorbability of PFCs Batch kinetic tests in UPW were conducted to determine the adsorbability of PFCs in the absence of NOM. Tests were conducted with 15 mg/l (S-)PAC in UPW at an ionic strength of 0.02 and ph 7.0. Kinetic test results for the ten (S-)PACs in UPW are presented in Figure

59 Figure 4.1 Effect of (S-)PAC type and particle size on PFC adsorption kinetics. 41

60 42

61 43

62 The data in Figure 4.1 illustrate that PFC removal increased with increasing carbon chain length, and that for a given chain length, sulfonates were more adsorbable than carboxylates. This result is consistent with previous findings by Hansen et al. (2010) who studied the removal of three sulfonate and five carboxylate PFCs in isotherms experiments dosed with 25 mg/l PAC. The longer the carbon-fluorine chain, the more hydrophobic the PFC, which increases its likelihood to accumulate at the water-carbon interface. For perfluorinated sulfonates, all of the carbon atoms are incorporated into the hydrophobic carbon-fluorine tail, as opposed to the perfluorinated carboxylates in which one of the carbon atoms is incorporated into the hydrophilic carboxylate acid functional group. This difference in carbon-flourine tail length explains the increased adsorbability of perfluorinated sulfonates compared to perfluorinated carboxylates with the same number of carbons. In addition, the results in Figure 4.1 illustrate that PFCs were not adsorbed as quickly with PACs as with S- PACs. This result can be explained by two factors: (1) the smaller particle size of S-PACs and (2) the larger mesopore volume of S-PACs (Table 4.1). The adsorption kinetics of micropollutants on PACs are typically controlled by the rate of intraparticle diffusion. The rate of such processes is proportional to the inverse of the square of the adsorbent diameter. Also, mesopores function as transport pores that aid PFCs diffusion into the porous activated carbon matrix. Regarding PFC adsorbability, removal of C4 was negligible with all (S-)PACs at the tested 15 mg/l carbon dose. Removal of C5 ranged from negligible to ~40% with S-C. Among the five S-PACs, S-C, a thermally activated wood based carbon, was the most effective 44

63 adsorbent for PFC removal with >90% uptake for eight of the ten PFCs after 15 minutes of contact time. Also interesting are results obtained with S-D, which suggest that smaller, more weakly adsorbed PFCs (e.g. C6, C7 and PFBS) were displaced by more strongly adsorbing PFCs such as C9, C10, and PFOS. To explore the potential for displacement further, a kinetic test was conducted in which C7 was spiked as the sole PFC into UPW. Figure 4.2 presents a comparison of C7 single-solute and PFC mixture results for S-D. Results in Figure 4.2 show that the single-solute data also exhibit a removal peak at a contact time of 5 minutes and a subsequent deterioration in removal, possibly due to the presence of trace levels of organic matter that was not removed in the treatment system that produces UPW from Raleigh tap water. Alternatively, if C7 removal took place on anion exchange sites on the activated carbon surface, chloride and/or bicarbonate anions in amended UPW could have displaced C7. C7 release with increasing contact time was less pronounced in the single-solute system than in the system containing the PFC mixture. Thus, the results shown in Figure 4.2 leave open the possibility that C7 was displaced by more strongly adsorbing PFCs when tested in the presence of other PFCs. 45

64 Figure 4.2 Comparison of C7 uptake rates from single-solute solution and PFC mixture. To evaluate the effect of NOM on the effectiveness of (S-)PAC for PFC removal, batch kinetic tests were conducted in CCR water for each of the five as-received PACs and their corresponding S-PACs. Figure 4.3 summarizes the PFC uptake rates obtained for each carbon. 46

65 Figure 4.3 Effect of activated carbon type and particle size on PFC uptake rates from CCR. 47

66 48

67 49

68 Consistent with previously presented results, PFC adsorbability increased with increasing carbon chain length, and sulfonates of a given carbon chain length were more adsorbable than carboxylates. A decrease in the adsorptive capacity of PFCs in CCR compared to UPW (Figure 4.1) was observed. Greater than 95% removal was achieved for more adsorbable PFCs in UPW with all (S-)PACs, but in CCR water the as-received PACs did not achieve more than 50% removal of any PFC and only two S-PACs (S-A and S-C) achieved more than 90% removal of C10 and PFOS, the two most adsorbable PFCs studied. CCR water contains NOM at a concentration of of 4.5 mg/l TOC, which is orders of magnitude higher than the ng/l PFC concentrations. Most likely, NOM competed with PFCs for adsorption sites and decreased the PFC uptake capacity of the carbons. One objective of the batch kinetic tests was to dermine if the blending of microporous and mesoporous (S-)PACs would be advantageous for PFC removal. The blending of carbons may be effective because of the wide range of molecular weights (MWs) represented by the ten PFCs selected for this research. However, batch kinetic test results indicate that carbons that were most effective for the removal of PFCs with higher MWs (e.g. carbons S-A and S- C in CCR, Figure 4.3) were also most effective for the removal of PFCs with lower MWs. For example, measureable removal of C7 from CCR could only be achieved with carbons S- A and S-C. Therefore, blending of carbons would not offer any advantages for PFC removal from CCR. 50

69 Effect of Background Water ph and Ionic Strength To evaluate the effect of solution ph and ionic strength on PFC adsorbability, batch kinetic tests for were conducted with (S-)C in UPW at ph 5.5, 7.0, and 8.5. Ionic strengths of and 0.02 were tested at ph 5.5 and 7.0. At ph 8.5, ionic strengths of 0.002, 0.01, and 0.02 were tested. The (S-)C carbon pair was selected because it was the most effective in batch kinetic tests evaluating PFC removal from UPW and CCR. Effects of solution ph and ionic strength on PFC uptake rates are summarized in Figure

70 Figure 4.4 Effect of ph and ionic strength on PFC removal by (S-)C 52

71 . 53

72 54

73 55

74 56

75 The effects of ionic strength and ph on PFC adsorbability are best illustrated by the results obtained with C5 because C5 removal percentages fell into a range in which differences in adsorptive uptake were measurable. In contrast, removal percentages for C7-C10, PFHS, and PFOS were too high and for C4 too low at the tested (S-)PAC dose of 15 mg/l to make meaningful comparisons. Results obtained with C5 suggest that C5 adsorbability increased with decreasing ph and ionic strength. For both carbons C and S-C, C5 removal was highest at ph 5.5 and an ionic strength of Increasing C5 removal with decreasing ph may be a result of one or both of the following factors: (1) the degree of C5 ionization decreased with decreasing ph, and (2) the net surface charge of C and S-C became more positive (ph pzc values are 10.7 and 7.9 for C and S-C, respectively) with decreasing ph. A smaller degree of ionization may increase adsorbability because a larger percentage of the adsorbate is present in the neutral form, which is less soluble and hence more adsorbable. However, this effect was likely small as the pka of C5 is approximately 0.5 (Rayne and Forest, 2009). The second effect, increasingly positive carbon pore surface charge with decreasing ph, could increase (1) the strength of the electrostatic interaction between the anionic C5 and the carbon surface and/or (2) the concentration of anion exchange sites on the carbon surface. The observation that C5 adsorbability at a low ionic strength of was somewhat higher than at the higher ionic strength of 0.02 at ph 5.5 is consistent with screening of attractive electrostatic forces between the positively charged carbon surface and the negatively charged C5 at the higher ionic strength. The screening explanation assumes that adsorption was the principal PFC removal mechanism. Alternatively, if anion exchange played an important role in PFC uptake 57

76 by activated carbon, then a higher concentration of competing anions at the higher ionic strength could explain lower PFC uptake at higher ionic strength. At ph 7.0, an increase in ionic strength lowered C5 adsorbability on C, but not S-C, a result that may be related to the higher ph pzc of the as-received PAC. A similar result was obtained at ph 8.5 except that no measurable C5 adsorption occurred on S-C at an ionic strength of 0.02, a result that could not be explained. For C6 and PFBS adsorption, a similar ph dependence was observed as for C5. The effect of ph on PFC adsorbability was also tested in the two drinking water sources. Batch kinetic tests were conducted with the C/S-C pair at the following conditions: (1) CCR at ambient ph 7.5, (2) CCR at ph 5.5, (3) CCR at ph 7.5 and adjusted with CaCl 2 to match the hardness of ORW, (4) ORW at ph 5.5, (5) ORW at ph 7.5, and (6) ORW following filtration through a 1-µm glass-fiber filter at ph 7.5. Results of the kinetic tests in natural water are summarized in Figure

77 Figure 4.5 Effect of ph, hardness, and turbidity in natural waters on PFC removal. 59

78 60

79 61

80 62

81 63

82 At all tested conditions, less than 30% removal of C4, C5, C6, and PFBS was achieved in the two drinking water sources. For the remaining PFCs, removals were more substantial and higher in ORW than in CCR. For example, after 2 hours of contact, C8 removal from ORW by S-C exceeded 75% at both ph values while it was less than 50% from CCR at both ph values. ORW had a TOC of approximately 3.9 mg/l compared to CCR s 4.5 mg/l. Similarly, the UV 254 absorbance of ORW was approximately compared to CCR s While differences in NOM concentration and characteristics between the CCR and ORW affected PFC removal, ph appeared to have little effect in both waters. For example, for the C8 data obtained with S-C, removals at ph 5.5 and 7.5 were within 8% of each other for ORW and within 4% of each other for CCR. A similar lack of ph dependence on PFC adsorption from natural waters was obtained for all other PFCs. It is possible that NOM adsorption modified the carbon pore surface chemistry such that the net surface charge of the NOM-coated carbon surface varied more weakly with ph than the net surface charge of the virgin carbon surface. Also, NOM adsorption tends to lower the ph pzc of activated carbon pore surfaces (Morris and Newcombe, 1993), suggesting that the net positive surface charge and thus electrostatic attraction between PFC anions and the carbon surface was lower in the presence of NOM. Alternatively, if anion exchange plays an important role in the uptake of PFCs by activated carbon, NOM, which is anionic in nature, may have occupied the majority of anion exchange sites at both ph 5.5 and 7.5 such that PFC removal as not measureable affected by solution ph. Over the duration of the 2-hour batch kinetic test, the higher concentration of inorganic anions in ORW relative to CCR (Table 3.1) did not appear to 64

83 adversely affect PFC removal. A different result was obtained in isotherm tests as discussed below. A kinetic test was also conducted in CCR water to which CaCl 2 was added such that the adjusted hardness matched that of the ORW. The objective of this test was to explore whether a possible association between calcium and PFCs could affect PFC adsorbability. As shown in Figure 4.5, addition of CaCl 2 to CCR water had no measurable effect on PFC removal. This result further supports that differences in NOM concentration and characteristics principally explain differences in PFC uptake from CCR and ORW. An additional batch kinetic test was conducted with ORW after filtration through a 1- m glass fiber membrane filter to evaluate whether a turbidity of ~70 NTU interfered with PFC removal by (S-)PAC. As shown in Figure 4.5, the presence of this relatively high level of turbidity had no measurable effect on PFC removal by (S-)PAC. For kinetic tests conducted in natural waters, UV 254 absorbance samples were taken at time 0 and 120 minutes. Results are presented in Figure

84 Figure 4.6 Comparison of UV 254 removal in CCR and ORW. Carbon Dose: 15 mg/l. Contact Time: 2 hours. For all kinetic tests conducted in CCR the percent removal of UV 254 absorbing NOM was 6-10% higher with S-C compared to C after 120 minutes. For kinetic tests conducted in ORW, the percent removal of UV 254 was 13-17% higher with S-C compared to C. Grinding the asreceived PAC C to a finer particle size and increasing the mesopore volume increased both NOM adsorption kinetics and NOM adsorption capacity (Dunn, 2011). (S-)PAC Isotherms Adsorption isotherm experiments were conducted to compare the equilibrium adsorption capacities of PFCs on carbons C and S-C. Because of the wide range in PFC adsorbabilities, 10 to 15 carbon doses were required to obtain sufficient data to construct adsorption 66

85 isotherms for each of the ten PFCs. From the adsorption isotherms, the Freundlich parameters for each PFC were obtained in amended UPW, CCR, and ORW. PFC adsorption isotherms obtained with carbons C and S-C are presented in Figures 4.7 and 4.8, respectively, for amended UPW, in Figures 4.9 and 4.10, respectively, for CCR, and in Figures 4.11 and 4.12, respectively, for ORW. 67

86 (a) (b) Figure 4.7 PFC adsorption isotherms obtained with as-received carbon C in UPW: (a) carboxylates, (b) sulfonates. 68

87 (a) (b) Figure 4.8 PFC adsorption isotherms obtained with superfine carbon S-C in UPW: (a) carboxylates, (b) sulfonates. 69

88 (a) (b) Figure 4.9 PFC adsorption isotherms obtained with as-received carbon C in CCR: (a) carboxylates, (b) sulfonates. 70

89 (a) (b) Figure 4.10 PFC adsorption isotherms obtained with superfine carbon S-C in CCR: (a) carboxylates, (b) sulfonates. 71

90 (a) (b) Figure 4.11 PFC adsorption isotherms obtained with as-received carbon C in ORW: (a) carboxylates, (b) sulfonates. 72

91 (a) (b) Figure 4.12 PFC adsorption isotherms obtained with superfine carbon S-C in ORW: (a) carboxylates, (b) sulfonates. As was observed in batch kinetic tests, adsorption isotherm results showed that PFC adsorbability increased with increasing carbon chain length and that sulfonates of a given carbon chain length are more adsorbable than the corresponding carboxylates. 73

92 Freundlich K values presented in Table 4.3 and the solid-phase equilibrium concentration q calculated for equilibrium concentrations 10 ng/l and 100 ng/l presented in Table 4.4 quantitatively illustrate this trend. In amended UPW, PFC adsorption capacities were higher for C than for S-C. This result was unexpected because PFC adsorption isotherms in CCR and ORW were similar for C and S-C. The latter result suggests that wet-milling of as-received PACs did not change the volume of pores in which PFCs adsorb. Such as result is consistent with prior findings that wet-milling of PAC does not change the adsorption capacity for micropollutants (Dunn 2011). A possible reason for the lower PFC adsorption capacity of S-C is that displacement of PFCs by Cl - and/or HCO - 3 may have occurred, a mechanism that may have also been the cause of the C7 displacement shown in Figure 4.2. Relative to amended UPW results, PFC uptake capacities were lower in CCR and ORW. For C6-C10 and the sulfonates, q 10 values calculated for carbon C from adsorption isotherm data in CCR and ORW were % of those obtained in amended UPW. Similarly, q 100 values in CCR and ORW were % of those obtained in amended UPW. Reductions in q 100 values were slightly larger than the reductions in q 10 values. In general, the reduction in PFC uptake capacity became more pronounced as PFC carbon chain length and adsorbability increased. 74

93 Table 4.3 Freundlich K and 1/n values for carbons C and S-C UPW CCR ORW PFC C S-C C S-C C S-C K* 1/n K 1/n K 1/n K 1/n K 1/n K 1/n C C C C C C C PFBS PFHS PFOS * Units: (ng/mg)(l/ng) 1/n - Could not be determined due to only one valid isotherm point 75

94 Table 4.4 Equilibrium solid phase concentrations calculated for equilibrium concentrations of 10 and 100 ng/l UPW CCR ORW PFC C S-C C S-C C S-C q* 10 q 100 q 10 q 100 q 10 q 100 q 10 q 100 q 10 q 100 q 10 q 100 C C C C C C , C , PFBS PFHS PFOS 522 1, * Units: ng/g - Could not be calculated due to only one valid isotherm point + Could not be calculated due to the negative 1/n value 76

95 Apart from comparing the adsorbability of PFCs for a given carbon and water, the adsorbability of each PFC was compared between carbons C and S-C in UPW, CCR, ORW. The isotherm comparisons for each PFC are shown in Figure

96 Figure 4.13 Comparison of PFC adsorption isotherms for (S-)PACs in UPW, CCR, and ORW. 78

97 79

98 80

99 As shown in Figure 4.13, the equilibrium PFC uptake capacity was similar for C and S-C in CCR and ORW. Thus, wet-milling of the as-received PAC did not measurably change its PFC adsorption capacity. However, as illustrated in the batch kinetic tests, PFC uptake kinetics were faster with superfine carbon than with as-received PAC. Thus, from a PFC removal perspective, decreasing PAC particle size enhances adsorption kinetics, but not the equilibrium uptake capacity. The results in Figure 4.13 illustrate that equilibrium uptake capacities were higher in UPW than in ORW and CCR. As expected, the presence of natural organic matter affects not only PFC removal kinetics but also the equilibrium PFC adsorption capacity. (S-) PAC Jar Tests The objective of the jar tests was to evaluate the removal of PFCs with (S-)PACs under coagulation conditions typically used by surface water treatment plants. Factors evaluated included (S-)PAC type, (S-)PAC dose, coagulation ph, and timing of (S-)PAC addition relative to the addition of alum. Effect of (S-)PAC Type and Dose PFC removal results obtained in jar tests conducted with carbons (S-)B and (S-)C are shown in Figure Carbon doses of 15 and 25 mg/l were tested for all four carbons. In addition, a carbon dose of 50 mg/l was tested for carbons C and S-C. 81

100 Figure 4.14 Effect of (S-)PAC type and dose on PFC removal in jar tests. 82

101 83

102 84

103 PFC removal with alum only exceeded 20% for C10 and PFOS, but was negligible for all other PFCs. In the presence of 50 mg/l S-C, PFC removal ranged from negligible for C4 and C5 to >90% for C9, C10, and PFOS. For each carbon type, the S-PAC performed better than the as-received form. For C6 and PFBS all carbons performed similarly. For the more adsorbable PFCs C8-C10, PFHS and PFOS, S-C and S-B were both more effective for PFC removal than as-received C and B. Between S-B and S-C, carbon S-C was more effective than S-B for C7, C8, and PFHS removal, but S-B and S-C performed similarly for C9 removal. For C10 and PFOS removal, S-B and S-C performed similarly at a dose of 15 mg/l and S-B outperformed S-C at a dose of 25 mg/l. The adsorbent contact time during jar tests was only 46 minutes, suggesting that adsorption kinetics played an important role in determining PFC uptake. The larger mesopore volume and smaller particle size of S-C and S- B compared to C and B, respectively, likely increased PFC adsorption kinetics and resulted in greater PFC removal with the superfine carbons during jar testing. In addition to PFC concentration, UV 254 absorbance and turbidity data were collected after settling. Figure 4.15 presents the UV 254 removal data. 85

104 Figure 4.15 Comparison of UV 254 removals for different (S-)PAC types and doses. Based on Figure 4.15, little change in the UV 254 removal was observed from alum only to the jars containing (S-)PAC. At the highest dose of 50 mg/l, S-C and C produced similar UV 254 removals. At a dose of 25 mg/l, a small difference can be observed in the removal achieved by different carbons. S-B (73%) performed slightly better than B (67%) and S-C (61%) performed slightly better than C (54%). Settled water turbidity results are compared for carbons (S-)B and (S-)C as a function of adsorbent dose in Figure

105 Figure 4.16 Effect of carbon type and dose on settled water turbidity. Alum dose: 55 mg/l. Coagulation ph: 6.2. Different drums of CCR were used for the (S-)C (drum 1) and (S-)B (drum 2) tests, so the initial turbidity was not to same. For both carbons B and C, a lower (S-)PAC dose resulted in 87

106 a higher settled water turbidity, suggesting that floc settleability improved with increasing (S-)PAC dose. After 10 minutes of settling, the settled water turbidity with S-B was ~0.7 NTU higher than with B. The finding of a higher settled water turbidity with S-PAC is consistent with results obtained by Deng (2010) and suggest that smaller carbon particles are not as effectively incorporated into settleable floc. Effect of Coagulation ph Jar tests were also conducted to evaluate the effect of coagulation ph on PFC removal. PFC removal percentages obtained at ph 5.5, 6.2, and 7.0 are summarized in Figure

107 Figure 4.17 Effect of coagulation ph on PFC removal. Alum dose: 55 mg/l. (S-)C dose: 15 mg/l. For all PFCs, the difference in PFC removal was <10% at the tested ph values. Thus, coagulation ph did not have a measureable effect on PFC removal in the ph range of Figure 4.18 presents UV 254 removal results for the different coagulation ph values. 89

108 Figure 4.18 Effect of coagulation ph on UV 254 removal. Alum dose: 55 mg/l. (S-)C dose: 15 mg/l. Without the addition of (S-)PAC, the highest UV 254 removals were achieved at ph 5.5 and 6.2. There was no measureable difference in UV 254 removal between S-C and C at both ph 5.5 and 6.2. At ph 7.0, S-C performed better than C; however, UV 254 removal with both C and (S-)C was significantly lower than at ph 5.5 and 6.2. In addition, the effect of coagulation ph on settled water turbidity was analyzed after 3.5 and 10 minutes of sedimentation. Results are presented in Figure

109 Figure 4.19 Comparison of the effect of coagulation ph on settled water turbidity. The settled water turbidity after 3.5 minutes of sedimentation varied greatly with ph. With S- C, the settled water turbidity ranged from 2.6 NTU at ph 6.2 to 7.3 NTU at ph 5.5. At ph 5.5 and 7.0, the settled water turbidity obtained in the presence of S-C was higher than that obtained with C and alum only. This trend was also observed after 10 minutes of sedimentation. At ph 6.2, after 3.5 and 10 minutes of sedimentation, differences in settled water turbidity were small between alum only, C, and S-C. Effect of Timing of (S-)PAC Addition Jar tests were also conducted to determine how PFC removal is affected by the timing of (S-)PAC addition relative to the addition of coagulant. In all experiments, alum served as the coagulant and was added at a dose of 55 mg/l. (S-)PAC was added (1) 5 minutes prior to alum, (2) together with alum, and (3) 9 minutes into flocculation. PFC removal results are presented in Figure

110 Figure 4.20 Effect of the timing of (S-)PAC addition on PFC removal. Alum dose: 55 mg/l. Coagulation ph: 6.2. (S-)C dose: 15 mg/l. 92

111 93

112 Negligible removals of C4, C5, C6 and PFBS were achieved in the jar tests evaluating timing of (S-)PAC addition. Consistent with batch kinetic tests, the percent removal of the more adsorbable PFCs C8, C9, C10, PFHS, and PFOS was higher with S-C than with C. Furthermore, results in Figure 4.20, show that the timing of (S-)PAC addition had no measureable effect on PFC removal. The effect of the timing of (S-)PAC addition on settled water UV 254 absorbance and turbidity was also analyzed (Figure 4.21 and 4.22, respectively). Figure 4.21 Effect of the timing of (S-)PAC addition on settled water UV 254 absorbance. Alum dose: 15 mg/l. Coagulation ph: 6.2. (S-)C dose: 15 mg/l. 94

113 Figure 4.22 Effect of the timing of addition of coagulant and (S-)PAC on settled water turbidity. As shown in Figure 4.21, UV 254 removal in the presence of alum only was 63%. In the presence of 15 mg/l C, UV 254 removal increased to 70% and to 78% in the presence of 15 mg/l S-C. Compared to (S-)PAC alone, the incremental UV 254 removal achieved with (S- )PAC in the presence of alum was smaller, suggesting that (1) NOM fractions that can be removed by coagulation and adsorption overlapped and/or (2) that the presence of alum floc interfered with NOM adsorption. Based on Figure 4.21, UV 254 removal was independent of the order in which coagulant and (S-)PAC were added. After both 3.5 and 10 minutes of settling, the settled water turbidity varied little when the timing of carbon C addition was varied; the lowest settled water turbidity (2.88 NTU after 3.5 min, 0.51 NTU after 10 min) was obtained when alum and C were added together (Figure 4.22). The high settled water turbidity obtained after 3.5 minutes with coagulant only during the evaluation of carbon C is likely an anomaly. The settled water turbidity after both 3.5 and 95

114 10 minutes of settling was higher with S-C than with C. After 10 minutes of settling, settled turbidity values in the presence of S-C were 2.1 NTU with alum only, 1.7 NTU when S-C was added 5 min. before alum, and 2.7 NTU when S-C and alum were added together. When S-C was added 9 minutes into flocculation, the turbidity after 10 minutes of settling was considerably higher at 7.3 NTU. The higher turbidity can be attributed to the alum floc having mostly formed by 9 minutes of flocculation, at which point S-C was not incorporated as effectively into it. (S-)PAC Dose Requirements for PFC Removal Both adsorption isotherm data and jar test results were used to determine (S-)PAC doses needed to achieve 90% and 50% PFC removal. Applying a simplified EBC model to the isotherm data collected, predictions were made for the carbon dose required to achieve 90% removal of each PFC. The Freundlich parameters C 2,0 and n 2 developed for the EBC are presented in Appendix D for each PFC. Using the Freundlich parameters in Tables 4.3, D.1, and D.2, Equation 3.2, the PFC fraction remaining at equilibrium was calculated as a function of carbon dose. For CCR, comparisons of EBC model results, experimental isotherm data, and experimental jar test data for carbons C and S-C are presented in Figure

115 Figure 4.23 Effect of (S-)PAC dose on PFC removal from CCR at equilibrium (isotherm results) and non-equilibrium (jar test results) conditions. 97

116 98

117 99

118 100

119 EBC model results general agreed well with the experimental isotherm data. Exceptions are C9 and PFBS results for S-C, primarily because the slope of the isotherms in amended UPW were smaller than those in CCR. A higher fraction of PFCs remained in the settled jar test water compared to isotherm results at the same (S-)PAC dose because jar tests were mixed for 36 minutes instead of 3 weeks, and equilibrium was not reached. For ORW, no jar test data were collected. Comparisons of EBC model results and experimental isotherm data obtained in ORW with carbons C and S-C are presented in Figure

120 Figure 4.24 Effect of (S-)PAC dose on PFC removal from ORW at equilibrium (isotherm results) and non-equilibrium (jar test results) conditions. 102

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122 104

123 105

124 Again, EBC model results agreed generally well with experimental isotherm data. Apart from the previously mentioned exception for C9 and PFBS with S-C, an additional challenge with the ORW data was that usable isotherm data was obtained at only S-C dose for C8 and PFHS. Additional carbon doses need to be evaluated in future work to assess the goodness of fit of the EBC model for C8 and PFHS. From the EBC model results, the C and S-C doses required to achieve 90% removal was calculated for each PFC in CCR and ORW (Table 4.5). Table 4.5 Carbon Doses Required to Achieve 90% PFC Removal from CCR and ORW CCR ORW C (mg/l) S-C (mg/l) C (mg/l) S-C (mg/l) C4 * * * * C C C C C C PFBS PFHS PFOS * Not feasible Dose predictions could not be made for C4 because the slope of the isotherms in amended UPW was greater than 1. As shown in Table 4.5, the carbon dose required to achieve 90% PFC removal at equilibrium decreased with increasing carbon chain length, and was lower for sulfonates than for the corresponding carboxylic acid. In general, dose predictions for 90% PFC removal were similar for S-C and C. In cases for which differences between C and S-C dose predictions were larger (>30%), no consistent trend was observed that would suggest that either C or S-C is more effective for PFC uptake at equilibrium. In other words, 106

125 wet-milling of as-received PACs did not alter the equilibrium uptake capacity of PFCs from CCR and ORW. Jar test data and isotherm model results were used to estimate the C and S-C doses required to achieve 50% removal of PFCs. Carbon doses required to achieve 50% removal for each PFC are compared for jar tests and isotherms in Table 4.6. Table 4.6 Carbon Dose Required to Achieve 50% Removal of C (mg/l) S-C (mg/l) Isotherm Jar Test Isotherm Jar Test C4 * > 50 * > 50 C5 89 > > 50 C6 42 > > 50 C C < 15 C < 15 C < 15 PFBS 34 > > 50 PFHS < 15 PFOS < 15 * Not feasible The isotherm model predictions in Table 4.6 represent the minimum carbon dose required to achieve 50% PFC removal as adsorption equilibrium was reached. In contract, jar test results represent a more realistic carbon dose requirement for shorter contact times typical of conventional surface water treatment. For PFOS, EBC model results suggest that 5 mg/l of as-received C is required for 50% removal while jar test results suggest 16 mg/l. The same trend was observed with all PFCs with carbon C and S-C, with the exception of C7, for which the EBC model predicted 24 mg/l S-C and the jar test results suggest 22 mg/l S-C is required for 50% removal. Differences in isotherm and jar test predictions are also illustrated 107

126 with Figure For C4-C6 and PFBS, negligible removal was achieved over the (S-)C dose range of mg/l. For the longer chain PFCs C8-C10, PFHS, and PFOS, Figure 4.23 shows that the jar test data became less similar to the isotherm results as PFC chain length increased. Differences between isotherm and jar test results were more pronounced for asreceived carbon C than for the superfine S-C. The latter two trends suggest that the intraparticle diffusivity of PFCs decreased with increasing PFC size and that the resulting adverse effect on adsorption kinetics could be lessened by decreasing the adsorbent particle size. Additional dose predictions were made to compare S-C doses required to achieve 90% PFC removal in isotherm and jar tests. (Table 4.7). Table 4.7 Carbon Doses Required to Achieve 90% PFC Removal from CCR in Isotherm and Jar Tests S-C (mg/l) Isotherm Jar Test C4 * > 50 C5 238 > 50 C6 97 > 50 C7 38 > 50 C8 26 > 50 C C PFBS 61 > 50 PFHS 26 > 50 PFOS * Not feasible The difference between isotherm and jar test results for longer chain PFCs are better illustrated with 90% removal predictions. For C9, C10, and PFOS, isotherm results suggest that S-C doses in the mg/l range will produce 90% removal. In contrast, jar test data illustrate that approximately 45 mg/l S-C would be needed to achieve 90% removal of C9, 108

127 C10, and PFOS at typical treatment conditions. A carbon dose of 45 mg/l is impractically high for most utilities. PFC REMOVAL BY ANION EXCHANGE In addition to (S-)PAC treatment, anion exchange was explored as an alternative treatment method for PFC control. The objectives of the anion exchange component of this research were to evaluate the effects of (1) resin type, (2) resin dose, and (3) background water constituents on PFC removal. Batch kinetic tests were conducted with three strong base anion (SBA) resins and one weak base anion (WBA) resin in UPW, CCR, and ORW (see Table 3.4). PFC removal results of experiments conducted in UPW are presented in Figure

128 Figure 4.25 PFC uptake kinetics from amended UPW for four anion exchange resins. 110

129 111

130 112

131 As with (S-)PAC, increasing PFC uptake by the resins correlated with increasing carbon chain length. In addition, sulfonic acids exhibited a higher affinity for the anion exchange resins than the corresponding carboxylic acids. Resin 2, a gel-type polystyrene SBA resin showed the highest PFC uptake capacity, with >85% removal after 120 minutes for all ten PFCs studied. The PFC uptake capacity of resin 3, a macroporous polystyrene SBA resin, was only slightly less, with greater than 90% removal for eight of the PFCs, 80% removal of C5, and 60% removal of C4. The PFC uptake kinetics of the gel-type resin 2 and macroporous resin 3 were similar, a result that is not consistent with previous studies, that showed faster uptake kinetics with macroporous resins (Deng et al. 2010). The macroporous polyacrylic SBA resin 1 showed the fastest PFC uptake kinetics among the three SBA resins; however, the PFC removal capacity of resin 1 was not as high as that of resins 2 and 3. A prior PFC removal study conducted with resins designed for non-potable water treatment applications attributed the effectiveness of polyacrylic SBA resins to the more hydrophilic properties of polyacrylic matrix which aids in the movement of water into the pores (Deng et al. 2010). Resin 4, a weak base anion (WBA) resin, performed more poorly with less than 65% removal for all PFCs at a dose of 10 ml/l in UPW (Figure 4.25). The poorer performance of the WBA resin was most likely related to the ph at which kinetic tests were conducted (ph 7); WBA resins perform more effectively at lower ph, at which tertiary amine sites become protonated. To evaluate background water matrix effects, additional experiments were conducted to compare anion exchange resin performance in CCR and ORW. PFC removal results obtained with the four resins in CCR and ORW are presented in Figure Each panel in Figure 113

132 4.26 compares the effectiveness of the four resins for removing a given PFC from either CCR or ORW. 114

133 Figure 4.26 Effectiveness of four anion exchange resins for PFC removal from two drinking water sources (CCR and ORW). 115

134 116

135 117

136 118

137 For CCR and ORW, results were consistent with UPW results in terms of the order of resin performance. Resin 4 had the lowest percent removal for all PFCs. In CCR, resin doses of 1, 5, and 10 ml/l were tested. While an increase in resin dose from 1 to 5 ml/l led to a substantial increase in the removal of all PFCs, less than 10% additional removal was achieved when the resin doses was increased from 5 to 10 ml/l. For compounds C4-C6, resin 2 achieved the highest removal percentages from CCR at all studied doses. For compounds C8-C10, PFHS, and PFOS resin 1 achieved the highest removal percentages while resin 1 and 2 performed similarly for C7 and PFBS. Resins 1 and 2 are both macroporous, but differ in that resin 1 is a polyacrylic and resin 2 a polystyrene resin. UV 254 removal percentage achieved with the four resins in CCR and ORW are presented in Figure 4.27 and 4.28, respectively. UV254 Removal Figure 4.27 UV 254 Removal with resins 1-4 in CCR 119

138 UV254 Removal Figure 4.28 UV 254 Removal with resins 1-4 in ORW At the lowest resin dose of 1 ml/l, resin 1 produced the highest UB254 removals in both CCR and ORW (66 and 48%, respectively). When 10 ml/l was added to CCR resins 1-4 achieved similar UV 254 removals, from 84% removal with resin 2 to 91% removal with resin 3. In ORW, resin doses of 10 ml/l produced from 70% removal (resin 2) to 86% removal (resin 3). To evaluate the effect of longer contact times on PFC uptake, resin contact times were extended to 24 hrs in experiments conducted with with resins 1-4 in ORW and with resins 1-3 in CCR. Results are presented in Figure

139 Figure 4.29 Effect of resin contact time on PFC removal from CCR and ORW. Resin dose: 5 ml/l. 121

140 122

141 123

142 124

143 The results in Figure 4.29 show that resin 1, in both CCR and ORW, more quickly achieved the 24 hour resin capacity compared to resins 2 and 3. For example, in CCR resin 1 achieved 88% of the 24 hr capacity for PFOS after 2 hours while resins 2 and 3 achieved only 49% and 36%, respectively. This result confirms that PFC uptake kinetics are more rapid with the polyacrylic resin that with the polystyrene resins. The same trend was observed in ORW as well. Figure 4.29 further illustrates that resin 4 was not very effective for PFC removal. Even after 24 hrs, PFOS and C10 removals did not exceed 24%, and no measureable removal of the other eight PFCs was obtained. To assess the effect of background water quality on PFC uptake by anion exchange resins, PFC removal results were compared in Figure 4.30 for the following waters: (1) amended UPW, (2) CCR at ambient ph 7.5, (3) CCR at ph 5.5, (4) CCR with 200 mg/l sulfate added, and (5) ORW. 125

144 Figure 4.30 Effect of background water matrix on PFC removal with resin

145 127

146 128

147 For all PFCs, higher removal percentages were achieved with resin 1 in CCR than in UPW. Removal percentages from ORW fell somewhere between those obtained in amended UPW and CCR for each PFC. The results obtained in CCR and ORW are interesting because it appears that the presence of NOM enhanced the removal of PFCs with anion exchange resin. In other words, for anion exchange, NOM had the opposite effect on PFC removal than it had for activated carbon. Results obtained with resins 2-4 were similar and are presented in Appendix E. It is not clearly understood why the presence of NOM would enhance the removal of PFCs by anion exchange resins. One hypothesis is that PFCs associated with NOM were pulled into the resin matrix. A second hypothesis is that coating the resin with NOM modified the resin surface such that its affinity for PFCs increased. To test the hypothesis, a preloading test was conducted with resin 1 in CCR. CCR was mixed with 5 ml/l of resin 1 for 24 hours prior to the dosing of PFCs. Results for this experiment are presented in Figure 4.31 alongside the original kinetic test results obtained without preloading the resin. 129

148 Figure 4.31 Comparison of PFC removals obtained with resin 1 following NOM preloading and with original resin 1. Water: CCR. If the NOM had been physically pulling the PFCs into the resin pores, it would be expected that preloading the resin would decrease PFC removal achieved by resin 1. Except for C4, decreased PFC removal was not observed, however. For C4, a 20% decrease in removal resulted as a consequence of NOM preloading. Therefor it is more likely that NOM altered the resin surface, making it more amenable for PFC removal. The reason for NOM improving the performance of the resin is not clearly understood and should be evaluated in future tests. Preliminary tests were conducted to preload resin 1 with three organic acids (benzoic acid, hexenoic acid, and heptanoic acid) that exhibit common NOM moieties (aromatic, unsaturated aliphatic, saturated aliphatic). However, resin 1 had a negligible affinity for the three organic acids, as, as a result, modification of the resin surface was not achieved. Figure 4.30 also illustrates the effects of ph and sulfate addition on PFC removal from CCR. The ph 5.5 experiment was conducted with a resin dose of 5 ml/l. With the exception of 130

149 C4, PFC removals at ph 5.5 and 7.5 were similar. For C4, removal was 7% lower at ph 5.5, a difference that falls within the analytical variability of the LC-MS/MS method. Therefore, the results in Figure 4.30 suggest ph has little effect on the effectiveness of ion exchange processes for PFC removal in the ph range studied. To probe the effect of sulfate on PFC removal by anion exchange, 200 mg/l of SO 2-4 was added to CCR (ph 7.5). The background sulfate concentration of CCR was less than 5 mg/l SO 2-4. Sulfate has a relatively high affinity for ion exchange resins and was expected to compete for ion exchange sites and thus adversely affect PFC removal. The results in Figure 4.30 confirm that adding sulfate to CCR decreased removal of all PFCs. In addition, it should be considered that ORW has a background sulfate concentration of 63 mg/l SO The removal achieved for each PFC in ORW was consistently between those achieved non-amended CCR and sulfate-amended CCR. The reason for PFC removal in ORW being less than in non-amended CCR could be attributed to (1) the lower NOM concentration in ORW did not modify the resin surface in the same manner as was achieved with CCR and/or (2) the higher sulfate concentration in ORW decreased the resin capacity for PFCs. Results from Figure 4.30 demonstrate that both NOM and sulfate affect PFC removal, while ph in the range of has little effect on PFC removal. Resin Dose Requirements for PFC Removal Based on PFC removal resultsobtaiend with resins 1-4 in CCR, resin doses required for 90% PFC removal were determined (Table 4.8). 131

150 Table 4.8 Resin Dose Requirements for 90% PFC Removal from CCR Resin 1 Resin 2 Resin 3 Resin 4 ml/l mg/l ml/l mg/l ml/l mg/l ml/l mg/l C4 > 10 > > 10 > 2750 C5 > 10 > > 10 > 2750 C > 10 > 2750 C > 10 > 2750 C > 10 > 2750 C > 10 > 2750 C > 10 > 2750 PFBS > 10 > 2750 PFHS > 10 > 2750 PFOS The volume-based resin doses required to achieve 90% PFC removal in CCR highlight that the polystyrene resins 2 and 3 were the most effective for PFC removal followed by the polyacrylic resin 1. On a mass-basis, resin 3, which has a lower density than the other resins, required the lowest doses to achieve 90% PFC removal. The results in Table 4.8 further illustrate that resin 4 was the least effective with > 10 ml/l required to achieve 90% removal for nine of ten PFCs in CCR. 132

151 CHAPTER 5: SUMMARY AND CONCLUSIONS The principal objective of this research was to assess the effectiveness of powdered activated carbon (PAC) adsorption and anion exchange processes for the removal of PFCs from drinking water sources. Experiments to evaluate PFC removal with powdered activated carbon were conducted with five as-received activated carbons prepared from different base materials (lignite, coconut shell, wood, and bituminous coal) and with either chemical or thermal activation processes. In addition, each of the as-received carbons was wet-milled to a finer diameter to produce five superfine-pacs. Anion exchange processes were evaluated with four anion exchange resins: gel-type and macroporous polystyrene SBA, macroporous polyacrylic SBA, and gel-type polyacrylic WBA. Experiments were conducted in three waters: buffered ultrapure water (UPW), Cane Creek reservoir water (CCR), and Ohio River water (ORW). PFC Removal with Activated Carbon Batch kinetic tests were conducted to determine the adsorbability of ten PFCs on all ten (S-)PACs. Experiments were conducted in UPW, CCR, and ORW to determine the effect of background organic matter on PFC removal. In addition, the effects of ph and ionic strength on PFC removal were evaluated in amended UPW. Furthermore, the effects of ph, calcium hardness, and turbidity on PFC removal were evaluated in CCR and ORW. Key findings were: 133

152 PFC adsorbability increased with increasing PFC carbon chain length. With all five S-PACs, removal of C4 was negligible while removal of C7-C10 from UPW was greater than 90%. As the hydrophobic carbon-fluorine tail of PFCs increases in length, PFCs accumulate more readily at the water-carbon interface. Adsorbability of sulfonates of a given carbon chain length was greater than that of the corresponding carboxylates. With all S-PACs, removal of C4 was negligible while removal of PFBS was greater than 75% in amended UPW. For carboxylates, one of the carbons is incorporated in the hydrophilic carboxylic acid head. In contrast, all of the carbons in a sulfonate reside within the hydrophobic carbon-fluorine tail. PFC adsorption kinetics were faster with S-PACs compared to as-received PACs. This result can be explained by two factors: (1) smaller particle size and (2) larger mesopore volume. Mesopores function as transport pores, aiding the movement of micropollutants into the carbon pore matrix. The presence of NOM decreased the effectiveness of activated carbon for PFC removal. Greater than 95% removal was achieved for more adsorbable PFCs in UPW with all ten (S-)PACs. In CCR water, as-received PACs did not achieve more than 55% removal of any PFC and only two S-PACs (S-A and S-C) yielded more than 90% removal of C10 and PFOS, the two most adsorbable PFCs studied. The blending of microporous and mesoporous carbons would not have been effective for increasing PFC removal because the carbons that were most effective for the removal of PFCs with higher molecular weights were also most effective for the removal of PFCs with lower molecular weights. The most effective carbons for PFC 134

153 removal (S-A and S-C) exhibited basin surface (ph PZC > 7), a large micropore volume, and a large mesopore volume. Experiments conducted with carbons C and S-C in UPW indicated that PFC removal increased with decreasing ph and ionic strength. The most likely explanation is that the decrease in ph increased the net positive carbon pore surface charge, increasing the electrostatic forces between the carbon and the anionic PFCs and/or the surface concentration of anion exchange sites. The decrease in adsorbability with increasing ionic strength may be attributable to the screening of attractive electrostatic forces between the PFC anions and positive carbon surface at higher ionic strengths and/or the displacement of PFCs by a higher concentration of inorganic anions competing for anion exchange sites. Experiments conducted with carbons C and S-C in CCR and ORW at ph 5.5 and 7.5 showed that ph did not affect PFC removal by activated carbon in natural waters. The surface charge of the NOM-coated carbon surface may have varied more weakly with ph than that of the virgin carbon surface. Alternatively, anionic NOM may have occupied the majority of the additional anion exchange sites that resulted as ph was decreased from 7.5 to 5.5. The addition of calcium hardness to CCR did not have a measureable effect on PFC removal by (S-)PAC. Filtering ORW water to remove turbidity (71 NTU) did not have a measureable effect on PFC removal by (S-)PAC. 135

154 Adsorption isotherm tests were conducted with C and S-C in amended UPW, CCR, and ORW to determine the equilibrium adsorption capacities of the carbons for PFCs and the effect of background water constituents on PFC adsorption capacity. Isotherm results confirmed the trend of increasing PFC adsorbability with increasing carbon chain length. In addition, isotherm results showed that sulfonates of a given carbon chain length are more adsorbable than the corresponding carboxylates. PFC equilibrium uptake capacities were higher in UPW than in CCR and ORW. In the presence of NOM, adsorption capacities of higher MW PFCs were generally most strongly affected. E.g., for as-received C is CCR, C10 and PFOS adsorption capacities at an equilibrium aqueous phase concentration of 10 ng/l were only about 4% of those obtained in amended UPW. PFC adsorption isotherms for C and S-C were similar in CCR and ORW. This result indicates that wet-milling of as-received PFCs did not affect the equilibrium uptake capacity for PFCs. Jar tests were conducted to evaluate PFC removal by (S-)PAC during alum coagulation. The following factors were evaluated in jar tests: (1) (S-)PAC type, (2) (S-)PAC dose, (3) coagulation ph, and (4) timing of (S-)PAC addition relative to the addition of alum. In tests evaluating carbon dose (15-50 mg/l) and type ((S-)C and (S-)B), PFC removal increased with (S-)PAC dose and, with respect to carbon type, carbon S-C was generally most effective. At an S-C dose of 50 mg/l, PFC removal ranged from negligible for C4 and C5 to >90% for C9, C10, and PFOS. 136

155 For carbons C and S-C, coagulation ph values of 5.5, 6.2. and 7.0 did not have a measureable effect on PFC removal from CCR. The timing of (S-)PAC addition relative to the addition of alum had no measureable effect on PFC removal from CCR. For all jar tests, application of S-C resulted in a higher settled water turbidity than C. This result suggests that the smaller S-PAC particles were not as effectively incorporated into settleable floc as the as-received PAC particles. Settled water turbidities were highest when S-PAC was added 9 minutes into flocculation. Both adsorption isotherm and jar test results were used to determine (S-)PAC doses needed to achieve 90% and 50% PFC removal. Isotherm model results suggest that >90% removal can be achieved for C8-C10, PFHS and PFOS when 25 mg/l C or S-C is added to CCR. Similar results were obtained with ORW. In comparison to isotherm model results, carbon dose estimates from jar tests were higher because adsorption equilibrium was not reached during jar testing. For example, isotherm results suggest that a S-C dose of 12 mg/l is required for 90% PFOS removal while jar test results suggest a dose of 42 mg/l. The dose estimates obtained with the isotherm model, constructed with data obtained at equilibrium conditions, represent the minimum carbon doses required for PFC removal. Jar test results represent more realistic doses for contact times more likely achieved in a conventional surface water treatment plants. 137

156 PFC Removal with Anion Exchange Resins Experiments were conducted with four anion exchange resins to determine the following effects on PFC removal: (1) resin type, (2) resin dose, and (3) background water constituents. Kinetic tests were conducted with resin doses of 1, 5, and 10 ml/l in amended UPW, CCR, and ORW. PFC affinity for anion exchange resins increased with increasing carbon chain length, and sulfonates of a given carbon chain length had a higher affinity than the corresponding carboxylate. In amended UPW, polystyrene SBA resins (resins 2 and 3) achieved the highest PFC removal. No measureable difference in PFC removal was achieved between gel-type and macroporous polystyrene SBA resin. In CCR, polystyrene SBA resins (resins 2 and 3) achieved the highest PFC removal at doses of 5 and 10 ml/l and for C4-C7 and PFBS at a resin dose of 1 ml/l. For C8- C10, PFHS and PFOS, the polyacrylic SBA resin 1 achieved the highest PFC removal at a dose of 1 ml/l. WBA resin (resin 4) was the least effective for PFC removal from amended UPW, CCR, and ORW. At ph 5.5 and 7.5, similar PFC removals were obtained by resin 1 in CCR. In the presence of NOM, higher PFC removals were obtained with all four anion exchange resins than in the absence of NOM. Results obtained with resins that were 138

157 preloaded with NOM indicate that NOM may alter the surface of the resin such that PFC affinity for the resin increased. The mechanism for this is not clearly understood and should be evaluated further. Additional preloading experiments with model compounds could be conducted to determine which NOM functionalities are responsible for the improved PFC removal. Higher PFC removal percentages were obtained in CCR, a higher TOC water, than in ORW. The addition of sulfate, an ion with a high affinity for anion exchange resins, decreased PFC uptake from CCR by resin 1. Higher sulfate concentration in ORW (64 mg/l) compared to CCR (4.6 mg/l), perhaps in conjunction with the lower NOM concentration, may explain the lower PFC removal achieved in ORW compared to CCR. Polystyrene SBA resins (resins 2 and 3) were able to achieve >90% removal of C4 and C5 at doses of 5 and 10 ml/l in CCR and ORW. Overall, the results of this research show that adsorptive removal of C4, C5, C6, C7, and PFBS from drinking water sources requires activated carbon doses that are too high to be practical. In contrast, polystyrene-based SBA exchange resins effectively removed all 10 PFCs from drinking water sources. Before the application of anion exchange for PFC removal can be recommended, effective resin regeneration strategies need to be developed. Alternatively, a hybrid adsorption/anion exchange treatment strategy may be most effective 139

158 for PFC removal, in which more strongly adsorbing PFCs are initially removed by activated carbon and the more weakly adsorbing PFCs subsequently by anion exchange. 140

159 REFERENCES Ahrens, L Polyfluoroalkyl compounds in the aquatic environment: a review of their occurrence and fate. Journal of Environmental Monitoring, 13: Bao, J., W. Liu, L. Liu, Y. Jin, J. Dai, X. Ran, Z. Zhang, and S. Tsuda Perfluorinated Compounds in the Environment and the Blood of Residents Living near Fluorochemical Plants in Fuxin, China. Environmental Science and Technology, 45(19): Betts, K.S Perfluoroalkyl Acids: What is the evidence telling us?. Environmental Health Perspectives, 115(5):A250-A256. Boulanger, B., J. Vargo, J.L. Schnoor, and K.C. Hornbuckle Detection of Perfluorooctane Surfactants in Great Lakes Water. Environmental Science and Technology, 38(15): Carter, K.E., and J. Farrell Removal of Perfluorooctane and Perfluorobutane Sulfonate from Water via Carbon Adsorption and Ion Exchange. Separation Science and Technology, 45(6): Clifford, D., Sorg, T.J., and G.L. Ghurye Ion Exchange and Adsorption of Inorganic Contaminants. In Water Quality and Treatment: A Handbook on Drinking Water. Edited by American Water Works Association and J.K Edzwald. McGraw Hill. Deng, S., Q. Yu, J. Huang, and G. Yu Removal of perfluorooctane sulfonate from wastewater by anion exchange resins: Effects of resin properties and solution chemistry. Water Research, 44(18): Dunn, S.E Effect of Powdered Activated Carbon Base Material and Size on Dinsinfection By-Product Precursor and Trace Organic Pollutant Removal. Master s thesis. North Carolina State University, Raleigh, NC. Emmett, E.A., S.F. Susan, Z. Hong Community exposure to perfluorooctanoate: Relationships between serum concentrations and exposure sources. Journal of Occupational and Environmental Medicine, 48(8): Fei, C., J.K. McLaughlin, R.E. Tarone, and J. Olsen Perfluorinated Chemicals and Fetal Growth: A Study within the Danish National Birth Cohort. Environmental Health Perspectives, 115(11): Giesy, J.P., and K. Kannan Perfluorochemical Surfactants in the Environment. Environmental Science and Technology, 36(7): 147A-152A. 141

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163 APPENDICES 145

164 Appendix A: Evaluation of PFC Loss to Bottles and Jars Figure A.1 Comparison of time 0 and 2 and 3 week PFC Concentrations in isotherm bottles. Figure A.2 Comparison of time 0 and 46 minute PFC Concentration in acrylic jar. 146

165 Appendix B: Representative LC-MS/MS Chromatograms Representative chromatograms for the carboxylates and sulfonates are shown in Figures A.1 and A.2, respectively. Representative chromatograms for the internal standards are shown in Figure A

166 Figure B.1 Chromatograms for carboxylates C4 C

167 C4 C5 C6 C7 C8 149

168 C9 C10 150

169 PFBS PFHS PFOS Figure B.2 Chromatograms for the sulfonates PFBS, PFHS, and PFOS. 151

170 Figure B.3 Chromatograms for internal standards. 152

171 MC4 MC6 MC8 MC10 MPFHxS MPFOS 153