|-f|^ United States Office of Water EPA 822P24003 Environmental Protection Office of Science and December 2024 Bail M % Agency Technology DRAFT Human Health Ambient Water Quality Criteria: Perfluorobutane Sulfonic Acid (PFBS) and Related Salts ------- Acknowledgements This document was prepared by the Health and Ecological Criteria Division, Office of Science and Technology, Office of Water (OW) of the U.S. Environmental Protection Agency (EPA). The OW scientists and managers who provided valuable contributions and direction in the development of these recommended water quality criteria are, from OST: Brandi Echols, PhD; Casey Lindberg, PhD; Czarina Cooper, MPH; Brittany Jacobs, PhD; Carlye Austin, PhD; Kelly Cunningham, MS (formerly OST); Erica Fleisig; Susan Euling, PhD; and Colleen Flaherty, MS; and, from the Office of Research and Development (ORD): Jason Lambert, PhD. ------- Table of Contents 1 Introduction: Background and Scope 1 2 Problem Formulation 2 2.1 Uses and Sources of PFBS 3 2.2 Environmental Fate and Transport in the Environment 4 2.3 Occurrence and Detection in Sources Relevant to Ambient Water Quality Criteria 4 2.3.1 Occurrence in Surface Water 5 2.3.2 Occurrence in Freshwater and Estuarine Fish and Shellfish 7 3 Criteria Formulas: Analysis Plan 7 4 AWQC Input Parameters 9 4.1 Exposure Factor Inputs 9 4.1.1 Body Weight 10 4.1.2 Drinking Water Intake Rate 10 4.1.3 Fish Consumption Rate 11 4.2 Bioaccumulation Factor (BAF) 11 4.2.1 Approach 11 4.2.2 Data Selection and Evaluation 13 4.2.3 BAFs for PFBS 15 5 Selection of Toxicity Value 17 5.1 Approach 17 5.2 Toxicity Value for PFBS 19 5.2.1 Reference Dose 19 5.2.2 Cancer Slope Factor 19 6 Relative Source Contribution (RSC) Derivation 20 6.1 Approach 20 6.2 Summary of Potential Exposure Sources of PFBS Other Than Water and Freshwater and Estuarine Fish/Shellfish 21 6.2.1 Dietary Sources 21 6.2.2 Food Contact Materials 23 6.2.3 Consumer Product Uses 24 6.2.4 Indoor Dust 26 6.2.5 Air 27 iii ------- 6.2.6 Soil 28 6.2.7 Summary and Recommended RSC for PFBS 29 7 Criteria Derivation: Analysis 32 7.1 AWQC for Noncarcinogenic Toxicological Effects 32 7.2 AWQC for Carcinogenic Toxicological Effects 33 7.3 AWQC Summary for PFBS 33 8 Consideration of Noncancer Health Risks from PFAS Mixtures 34 9 Chemical Name and Synonyms 35 10 References 36 Appendix A: Summary of Supporting Literature for Surface Water Occurrence 56 Appendix B: Bioaccumulation Factor (BAF) Supporting Information 68 Appendix C: Supporting Literature for Deriving the Relative Source Contribution 73 Appendix D: Comparative Analysis for Potentially Sensitive Populations for PFBS 89 iv ------- 1 Introduction: Background and Scope The U.S. Environmental Protection Agency's national recommended ambient water quality criteria (AWQC) for human health are scientifically derived numeric values that define ambient water concentrations that are expected to protect human health from the adverse effects of individual pollutants in ambient water. Section 304(a)(1) of the Clean Water Act (CWA) requires the EPA to develop and publish, and from time-to-time revise, recommended criteria for the protection of water quality that accurately reflect the latest scientific knowledge. Water quality criteria for human health developed under section 304(a) are based solely on data and scientific judgments about the relationship between pollutant concentrations and human health effects. Section 304(a) criteria do not reflect consideration of economic impacts or the technological feasibility of meeting pollutant concentrations in ambient water. The EPA's recommended section 304(a) criteria provide technical information for states and authorized Tribes3 to consider and use in adopting water quality standards that ultimately provide the basis for assessing water body health and controlling discharges of pollutants into waters of the United States. Under the CWA and its implementing regulations, states and authorized Tribes are required to adopt water quality criteria to protect the designated uses of waters (e.g., public water supply, aquatic life, recreational use, industrial use). The EPA's recommended water quality criteria do not substitute for the CWA or regulations, nor are they regulations themselves. Thus, the EPA's recommended criteria do not establish legal rights or obligations or impose legally binding requirements and are not final agency actions. States and authorized Tribes may adopt, where appropriate, other scientifically defensible water quality criteria that differ from these recommendations. The EPA's water quality standards regulation at 40 CFR 131.20(a) requires states and authorized Tribes to consider any new or updated national section 304(a) recommended criteria as part of their triennial review process, and, if the state or authorized Tribe does not adopt new or revised criteria for parameters that correspond to those new or revised 304(a) criteria, to provide an explanation when it submits its triennial review to EPA. This requirement is to ensure that state or Tribal water quality standards reflect the current science and protect applicable designated uses. The water quality criteria that are the subject of this document are draft national AWQC recommendations for human health issued under CWA section 304(a). Unless expressly indicated otherwise, all references to "human health criteria," "criteria," "water quality criteria," "ambient water quality criteria recommendations," or similar variants thereof are references to draft national AWQC recommendations for human health. a Throughout this document, the term states means the 50 states, the District of Columbia, the Commonwealth of Puerto Rico, the Virgin Islands, Guam, American Samoa, and the Commonwealth of the Northern Mariana Islands. The term authorized Tribe or Tribe means an Indian Tribe authorized for treatment in a manner similar to a state under CWA section 518 for the purposes of section 303(c) water quality standards. ------- Perfluorobutane sulfonic acid (PFBS) is a member of the per- and polyfluoroalkyl substances (PFAS) class. PFAS are a large class of thousands of synthetic chemicals that have been in use in the United States and around the world since the 1940s (EPA, 2018). The ability for PFAS to withstand heat and repel water and stains makes them useful in a wide variety of consumer, commercial, and industrial products, and in the manufacturing of other products and chemicals. The current scientific evidence has shown the potential for harmful health effects after human exposure to certain PFAS. The persistence and resistance to hydrolysis, photolysis, metabolism, and microbial degradation of PFAS raise additional concerns about long-term exposure and human health effects. The EPA developed the draft human health criteria (HHC) for PFBS to reflect the latest scientific information for input values, including exposure factors (i.e., body weight [BW], drinking water intake [DWI] rate, and fish consumption rate [FCR]), bioaccumulation factors (BAFs), human health toxicity values (i.e., reference dose [RfD]), and relative source contribution (RSC). The draft criteria are based on the EPA's current Methodology for Deriving Ambient Water Quality Criteria for the Protection of Human Health (2000a), which is referred to as the "2000 Methodology" in this document (EPA, 2000a). 2 Problem Formulation Problem formulation provides a strategic framework for ambient water quality criteria development to systematically identify the major factors and chemical-specific scientific issues to be considered in the assessment (EPA, 2014a). The structure of this draft criteria document is intended to be consistent with general concepts of health assessments as described in the EPA's Framework for Human Health Risk Assessment to Inform Decision Making (EPA, 2014a). In developing AWQC, the EPA follows the assessment method outlined in the 2000 Methodology (EPA, 2000a). The 2000 Methodology describes different approaches for addressing water and nonwater exposure pathways to derive human health AWQC depending on the toxicological endpoint of concern, the toxicological effect (noncarcinogenic or carcinogenic), and whether toxicity is considered a linear or threshold effect. Water sources of human exposure include both consuming drinking water and eating fish or shellfish from inland and nearshore water bodies that have been contaminated with pollutants. For pollutants that exhibit a threshold of exposure below which deleterious human health effects are unlikely to occur, as is the case for noncarcinogens and nonlinear carcinogens, the EPA applies an RSC. The RSC is the percentage of the total exposure to a contaminant that is attributed to the combination of drinking water and eating freshwater and estuarine fish and shellfish, where the remainder of exposure is allocated to other sources of oral exposure and other routes of exposure. The RSC is calculated by examining the data for other sources (e.g., air, food, soil) and pathways of exposure following the exposure decision tree for calculation of an RSC described in the 2000 Methodology (EPA, 2000a). For carcinogenic substances for which the cancer slope factor was quantified using linear low- dose extrapolation, only the exposures from drinking water and fish ingestion are reflected in the human health AWQC; nonwater sources are not explicitly included, and no RSC is applied 2 ------- (EPA, 2000a). This is because in these situations, AWQC are derived with respect to the incremental lifetime cancer risk posed by the presence of a substance in ambient water, rather than an individual's total risk from all exposure sources. Therefore, the resulting AWQC represents the ambient water concentration that is expected to increase an individual's lifetime risk of cancer from exposure to the pollutant by no more than one chance in one million (10~6) for the general population (male and female adults, 21 years and older; referred to as "general population" herein), regardless of the additional lifetime cancer risk due to exposure, if any, to that substance from other sources. The EPA calculates AWQC at a 10"6 cancer risk level for the general population (EPA, 2000a). The 2000 Methodology recommends that states set human health criteria cancer risk levels for the target general population at either 10"5 or 10"6and also notes that states and authorized Tribes can choose a more stringent risk level, such as 10"7. For substances that are carcinogenic, the EPA takes an integrated approach by considering both cancer and noncancer effects when deriving AWQC (EPA, 2000a,b). Where sufficient data are available, the EPA first derives separate AWQC for both carcinogenic and noncarcinogenic toxicity endpoints and then selects the lower (more health protective) of the two values for the recommended AWQC. PFBS may exist in multiple forms, such as isomers or associated salts and each form may have a separate Chemical Abstracts Service registry number (CASRN) or no CASRN at all. Additionally, these compounds have various names under different classification systems. PFBS and its related salts are members of the group of PFAS known as short-chain perfluoroalkane sulfonates. PFBS is an acid that is generally present as the sulfonate anion at typical environmental pH values. Therefore, the conclusions in this document apply to all isomers of PFBS, as well as nonmetal salts of PFBS that would be expected to dissociate in aqueous solutions of pH ranging from 4 to 9. For purposes of this assessment, "PFBS" will signify the ion, acid, or any nonmetal salt of PFBS. 2.1 Uses and Sources of PFBS PFAS are manufactured chemicals that have been widely used in industrial and consumer processes and products over the past several decades in the United States due to their repellant and surfactant properties. PFAS are persistent chemicals based on their physicochemical properties. Concerns about persistence of PFAS stem from the resistance of these compounds to hydrolysis, photolysis, metabolism, and microbial degradation. PFBS has been used as a replacement chemical for perfluorooctane sulfonic acid (PFOS), a chemical that was voluntarily phased out (with some exceptions) by its primary U.S. manufacturer, 3M Company, by 2002 (3M, 2002; EPA, 2007). Prior to its use as a PFOS replacement, PFBS had been produced as a byproduct and was present in consumer products as an impurity (AECOM, 2019). Concerns arising in the early 2000s about the environmental persistence, bioaccumulation potential, and long half-lives of PFOS and other long-chained PFAS in humans resulted in their replacement with shorter-chain PFAS, such as PFBS, in consumer products and applications (EPA, 2021a,b). PFBS and other shorter-chain PFAS possess 3 ------- the desired chemical properties of longer-chain PFAS, but have shorter half-lives in humans (EPA, 2021a,b). Environmental releases of PFBS may result directly from the production and use of PFBS itself, production and use of PFBS-related substances for various applications, and/or from the degradation of PFBS precursors (i.e., substances that may form PFBS during use, as a waste, or in the environment). PFBS is used in the manufacture of paints, cleaning agents, and water- and stain-repellent products and coatings (EPA, 2021a,b). PFBS has also been used as a mist suppressant for chrome electroplating and has been found associated with the use of aqueous film-forming foam (AFFF) (EPA, 2021a,b). PFBS has been detected in dust, carpeting and carpet cleaners, floor wax, and food packaging (ATSDR, 2021; EPA, 2021a,b). 2.2 Environmental Fate and Transport in the Environment The European Chemical Agency (ECHA) reports that PFBS is stable to hydrolysis, oxidation, and photodegradation in the atmosphere, and there have been no reports of abiotic degradation under environmental conditions (ECHA, 2019). The persistence of PFBS has been attributed to the strong carbon-fluorine (C-F) bond. PFBS has a high solubility in water (52.6 grams per liter [g/L] at 22.5-24 degrees Celsius (°C) for the potassium salt) and high mobility in the environment15 (log Koc 1.2 to 2.7) (ECHA, 2019). The Norwegian Environment Agency conducted a literature review of physicochemical properties and environmental monitoring data for PFBS to assist an evaluation under Registration, Evaluation, Authorization and Restriction of Chemicals (Arp and Slinde, 2018). No studies were identified that observed degradation of PFBS under environmental conditions, including atmospheric photolysis. The review determined that the air-water partition coefficient (Kaw) for PFBS is too low to measure and that volatilization from water is negligible, but that the presence of PFBS in ambient air can result from direct emissions or transport of droplets in contaminated water. ECHA (2019) modeled photodegradation of PFBS in air and concluded that PFBS has the potential for long-range transport. 2.3 Occurrence and Detection in Sources Relevant to Ambient Water Quality Criteria PFBS has been detected in a variety of environmental matrices. Studies describing the occurrence and detection of PFBS in sources relevant to ambient water quality criteria, including ambient water, fish, and shellfish, were identified through systematic literature searches of the peer reviewed and gray literature (see Section 6.2 below and Appendix B of EPA, 2024a for additional detail) and are described below. Additional occurrence information for sources other than ambient water (e.g., air, food, soil) is summarized in Section 6.2 as part of the determination of the RSC. b A measure of mobility is the sediment or soil organic carbon-water partition coefficient (Koc) with units of liters per kilogram (L/kg) and commonly expressed as log Koc, which is unitless. 4 ------- 2.3.1 Occurrence in Surface Water Studies evaluating the occurrence of PFBS in surface water in North America or Europe are summarized in Table A-l. Broadly, studies either targeted surface waters used as drinking water sources, surface waters known to be contaminated with PFAS (as reported by the study authors), or surface waters over a relatively large geographic area (i.e., statewide) with some or no known point sources of PFAS. Zhang et al. (2016) identified major sources of surface water PFAS contamination by collecting samples from 37 rivers and estuaries in the northeastern United States (metropolitan New York area and Rhode Island). PFBS was detected at 82% of sites and the range of PFBS concentrations was nondetect (ND) to 6.2 nanograms per liter (ng/L). Appleman et al. (2014) collected samples of surface water that were impacted by wastewater effluent discharge in several states. PFBS was detected in 64% of samples from 11 sites with concentrations ranging from ND to 47 ng/L. Several other studies from North America (four from the United States and two from Canada) evaluated surface waters from sites for which authors did not indicate whether sites were associated with any specific, known PFAS releases (Nakayama et al., 2010; Pan et al., 2018; Subedi et al., 2015; Veillette et al., 2012; Yeung et al., 2017). Nakayama et al. (2010) also collected samples across several states, but no specific source of PFAS was identified. The detection frequency (DF) of PFBS in the Nakayama et al. (2010) study was 43% with median and maximum levels of 0.71 ng/L and 84.1 ng/L, respectively. As reported in EPA (2024b), Pan et al. (2018) sampled surface water sites in the Delaware River with 100% DF, though PFBS levels were relatively low (0.52 ng/L to 4.20 ng/L); Yeung et al. (2017) reported results for a creek (PFBS concentration of 0.02 ng/L) and a river (no PFBS detected) in Canada. Veillette et al. (2012) analyzed surface water from an Arctic lake and detected PFBS at concentrations ranging from 0.011 ng/L to 0.024 ng/L. Subedi et al. (2015) evaluated lake water potentially impacted by septic effluent from adjacent residential properties, and detected PFBS in only one sample at a concentration of 0.26 ng/L. Additional available studies assessed surface water samples at U.S. sites contaminated with PFAS from nearby PFAS manufacturing facilities (ATSDR, 2021; Galloway et al., 2020; Newsted et al., 2017; Newton et al., 2017) or facilities that manufacture products containing PFAS (Lasier et al., 2011; Procopio et al., 2017; Zhang et al., 2016). A few of these studies identified potential point sources of PFAS contamination, including industrial facilities (e.g., textile mills, metal plating/coating facilities), airports, landfills, and wastewater treatment plants (WWTPs) (Galloway et al., 2020; Zhang et al., 2016). Among these sites, PFBS DFs (0% to 100%) and PFBS levels (ND to 336 ng/L) varied. In general, PFBS DFs that ranged from 0% to 3% were associated with samples collected upstream of PFAS point sources, and higher PFBS DFs (up to 100%) and PFBS concentrations were associated with samples collected downstream of point sources. An additional study (Lindstrom et al., 2011) sampled pond and stream surface water in areas impacted by up to 12 years of field applications of biosolids contaminated by a fluoropolymer manufacturer, and the maximum and mean PFBS concentrations were 208 ng/L and 26.3 ng/L, respectively. 5 ------- Another group of studies from the United States evaluated sites known to be contaminated from military installations with known or presumed AFFF use (Anderson et al., 2016; Nakayama et al., 2007; Post et al., 2013). The highest PFBS levels in ambient water reported among these available studies were from Anderson et al. (2016) who performed a national study of 40 AFFF- impacted sites across 10 military installations and reported a maximum PFBS concentration of 317,000 ng/L. Lescord et al. (2015) examined PFAS levels in Meretta Lake, a Canadian lake contaminated with runoff from an airport and military base, which are likely sources of PFAS from AFFF use. The authors reported a 70-fold greater mean PFBS concentration for the contaminated lake versus a control lake. In addition to AFFF, Nakayama et al. (2007) identified industrial sources, including metal-plating facilities and textile and paper production, as contributing to the total PFAS contamination in North Carolina's Cape Fear River Basin. Nakayama et al. (2007) reported a PFBS DF of 17% and PFBS concentrations ranging from ND to 9.41 ng/L at these sites. The EPA identified additional studies evaluating surface water samples from sites in Europe with known or suspected PFAS releases associated with AFFF use (Dauchy et al., 2017; Gobelius et al., 2018; Mussabek et al., 2019) or fluorochemical manufacturing (Bach et al., 2017; Boiteux et al., 2017; Gebbink et al., 2017; Valsecchi et al., 2015). PFBS levels were comparable at the AFFF-impacted sites (< 300 ng/L overall). Of the four study sites potentially contaminated based on proximity to fluorochemical manufacturing sites, two (from studies conducted in France) did not have PFBS detections (Bach et al., 2017; Boiteux et al., 2017). PFBS levels were low at most sampling locations of the remaining two studies (up to approximately 30 ng/L) except for the site in River Brenta in Italy (maximum PFBS concentration of 1,666 ng/L) which is also impacted by nearby textile and tannery manufacturers (Valsecchi et al., 2015). Eight studies in Europe evaluated areas close to urban areas, commercial activities, or industrial activities (e.g., textile manufacturing) (Boiteux et al., 2012; Eschauzier et al., 2012; Lorenzo et al., 2015; Rostkowski et al., 2009; Zhao et al., 2015) and/or wastewater effluent discharges (Labadie and Chevreuil, 2011; Lorenzo et al., 2015; Moller et al., 2010; Wilkinson et al., 2017). Among these sites, PFBS DFs varied (0 to 100%) and PFBS levels were less than 250 ng/L overall. Ten studies conducted in Europe evaluated sites with no known fluorochemical source of contamination (Ahrens et al., 2009a,b; Barreca et al., 2020; Ericson et al., 2008; Eriksson et al., 2013; Loos et al., 2017; Munoz et al., 2016; Pan et al., 2018; Shafique et al., 2017; Wagner et al., 2013). Pan et al. (2018) analyzed surface water from sites in the United Kingdom (Thames River), Germany and the Netherlands (Rhine River), and Sweden (Malaren Lake). While none of the sites sampled were proximate to known sources of PFAS, but PFBS was detected in all three water bodies. Concentrations of PFBS ranged from 0.46 ng/L to 146 ng/L; the highest level (146 ng/L) was detected in the Rhine River and was more than 20 times greater than any maximum level found in the other water bodies. In the remaining nine studies, reported PFBS levels ranged from ND to 26 ng/L, except for one study in Italy that reported a PFBS DF of 39% and levels in the |ag/L range at three out of 52 locations within the same river basin: Legnano (16,000 ng/L), Rho (15,000 ng/L), and Pero (3,400 ng/L) (Barreca et al., 2020). 6 ------- 2.3.2 Occurrence in Freshwater and Estuarine Fish and Shellfish Based on the available data collected to date, PFBS has been rarely detected in freshwater and estuarine fish and shellfish in the U.S. Several large-scale sampling efforts have been conducted by the EPA and other agencies to determine PFAS levels in fish. In the EPA's 2013-2014 National Rivers and Streams Assessment (NRSA), PFBS was detected at concentrations above the method detection limit (MDL) (0.1 ng/g) but below the quantitation limit (1 ng/g), at 0.571 ng/g in a largemouth bass fish fillet sample collected from Big Black River, Mississippi; 0.475 ng/g in a smallmouth bass fillet composite collected from Connecticut River, New Hampshire; and 0.148 ng/g in a walleye fillet composite collected from Chenango River, New York (EPA, 2020). However, in the 2008-2009 NRSA, PFBS was not detected in any fish species sampled (Stahl et al., 2014). In the EPA's 2015 Great Lakes Human Health Fish Fillet Tissue Study, PFBS was detected at a concentration of 0.36 ng/g in a smallmouth bass fillet composite collected from Lake Erie, New York (EPA, 2021c). In the National Rivers and Streams Assessment 2018-2019 (EPA, 2023) PFBS was a target chemical but was not detected in any of the fish samples analyzed. Note that PFBS was not a target chemical in the EPA's National Lake Fish Tissue Study (EPA, 2009) or the EPA's National Lakes Assessment 2017 (EPA, 2022a). PFBS was a target chemical for the National Lakes Assessment 2022 (EPA, 2024c), but was not detected in any of the fish samples analyzed (MDL 0.090 wet weight [ww]). More recently, PFBS has been detected in several estuarine species, including Irish pompano, silver porgy, grey snapper, and eastern oyster from the St. Lucie Estuary in the National Oceanic and Atmospheric Administration's (NOAA's) National Centers for Coastal Ocean Science, National Status and Trends Data (NOAA, 2024). 3 Criteria Formulas: Analysis Plan Human health AWQC for toxic pollutants may be necessary to protect designated uses of water bodies related to ingestion of water (i.e., public water supply or source water protection) and ingestion of freshwater/estuarine fish and shellfish. See CWA 303(c)(2)(A)-(B). Although the AWQC are based on chronic health effects data (both cancer and noncancer effects), the criteria are intended to also be protective against adverse effects that may reasonably be expected to occur as a result of elevated acute or short-term exposures (EPA, 2000a). Human health AWQC are expected to provide adequate protection not only for the general population over a lifetime of exposure, but also for sensitive life stages and subpopulations who, because of high water- or fish intake rates, or because of biological sensitivities, have an increased risk of receiving a dose that would elicit adverse effect (EPA, 2000a). The derivation of human health AWQC requires information about both the toxicological endpoints of concern from exposure to water pollutants and human exposure pathways for those pollutants. The EPA considers two primary pathways of human exposure to pollutants present in a particular water body when deriving human health 304(a) AWQC: (1) direct ingestion of drinking water obtained from the water body; and (2) consumption of fish and shellfish obtained from the water body. 7 ------- The equations for deriving human health AWQC are presented as Equations (Eqs.) 1 and 2 for noncancer and non-linear carcinogenic effects, and Eqs. 3 and 4 for linear carcinogenic effects. The EPA derives two separate recommended human health AWQC based on 1) the consumption of both water and aquatic organisms (Eq. 1), called "water + organism"; and 2) the consumption of freshwater/estuarine fish and shellfish alone (Eq. 2), called "organism only." The use of one criterion over the other depends on the designated use of a particular water body or water bodies (i.e., drinking water source and/or fishable waters). The EPA recommends applying organism only AWQC (Eq. 2) to a water body where the designated use includes supporting fishable uses under section 101(a) of the CWA but the water body is not a drinking water supply source (e.g., non-potable estuarine waters that support fish or shellfish for human consumption) (EPA, 2000a). The EPA recommends including the drinking water exposure pathway for ambient surface waters where drinking water is a designated use for the following reasons: (1) drinking water is a designated use for surface waters under the CWA, and therefore, criteria are needed to ensure that this designated use can be protected and maintained; (2) although they are rare, some public water supplies provide drinking water from surface water sources without treatment; (3) even among the majority of water supplies that do treat surface waters, existing treatments might not be effective for reducing levels of particular contaminants; and (4) in consideration of the agency's goals of pollution prevention, ambient waters should not be contaminated to a level where the burden of achieving health objectives is shifted away from those responsible for pollutant discharges and placed on downstream users that must bear the costs of upgraded or supplemental water treatment (EPA, 2000a). The equations for deriving the criteria values are as follows (EPA, 2000a): Equations for Noncancer and Nonlinear Carcinogen HHC: Consumption of water and organisms: AWQC = RfD x RSC x BW x 1.0QQC (Eq. 1) DWI + £f=2 (FCRi x BAFi) For consumption of organisms only: AWQC = RfD x RSC x BW x l.QQQc (Eq. 2) Z?=2 (FCRi x BAFi) Where: AWQC = ambient water quality criteria, expressed in micrograms per liter (|_ig/L) RfD = reference dose, expressed in milligrams per kilogram-day (mg/kg-d) RSC = relative source contribution, unitless BW = body weight, expressed in kg c 1,000 ng/mg is used to convert the units of mass from milligrams to micrograms. 8 ------- DWI = drinking water intake, expressed in L/d 2j!2 = summation of values for aquatic trophic levels (TLs), where the letter /' stands for the TLs to be considered, starting with TL 2 and proceeding to TL 4 FCRi = fish consumption rate for aquatic TLs (i) 2, 3, and 4, expressed in kg/d BAFi = bioaccumulation factor for aquatic TLs (i) 2, 3, and 4, expressed in L/kg Equations for Linear Carcinogens HHC: Consumption of water and organisms: AWQC = RSD x BW x l,000d (Eq. 3) DWI +Z?=2(FCRi x BAFi) For consumption of organisms only: AWQC = RSD x BW x 1.0QQd (Eq. 4) Yj\=2 (FCRi x BAFi) Where: AWQC = ambient water quality criteria, expressed in micrograms per liter (|_ig/L) RSD = RSD = risk specific dose; the cancer risk level (i.e., a target risk for the population; 1 in 1 million or 10"6) divided by the cancer slope factor (i.e., incidence of cancer relative to dose in units of [mg/kg/day]"1), expressed in milligrams per kilogram-day (mg/kg-d) BW = body weight, expressed in kg DWI = drinking water intake, expressed in L/d 24 = summation of values for aquatic trophic levels (TLs), where the letter /' stands for the TLs to be considered, starting with TL 2 and proceeding to TL 4 FCRi = fish consumption rate for aquatic TLs (i) 2, 3, and 4, expressed in kg/d BAFi = bioaccumulation factor for aquatic TLs (i) 2, 3, and 4, expressed in L/kg The EPA rounds AWQC to the number of significant figures in the least precise parameter as described in the 2000 Methodology (EPA, 2000a, Section 2.7.3). The EPA used a rounding procedure that is consistent with the 2000 Methodology (EPA, 2000a) and the 2015 HHC update (https://www.epa.gov/wqc/human-health-water-quality-criteria-and-methods-toxics). 4 AWQC Input Parameters 4.1 Exposure Factor Inputs National recommended HHC establish ambient concentrations of pollutants in waters of the United States which, if not exceeded, will protect the general population from adverse health impacts from those pollutants due to consumption of aquatic organisms (i.e., freshwater and estuarine fish and shellfish) and water (EPA, 2000a). It is the EPA's longstanding practice to set national recommended HHC at a level intended to be adequately protective of human exposure over a lifetime (EPA, 2000a). To accomplish this, the EPA uses a combination of median values, d 1,000 ng/mg is used to convert the units of mass from milligrams to micrograms. 9 ------- mean values, and percentile estimates for the HHC inputs consistent with the EPA's 2000 Methodology (EPA, 2000a). The EPA's assumptions afford an overall level of protection targeted at the high end of the general adult population (i.e., the target population or the criteria-basis population) (EPA, 2000a). This approach is reasonably conservative and appropriate to meet the goals of the CWA and the 304(a) criteria program (EPA, 2000a). If the EPA determines that another population or life stage (e.g., pregnant women and their fetuses, young children) is the target population, then exposure parameters for that target population or life stage could be considered in the derivation of the criteria (EPA, 2000a). Potentially sensitive life stages for PFBS are explored further in a comparative analysis in Appendix D. 4.1.1 Body Weight The BW for the general adult population including males and females, ages 21 years and older, was selected for the PFBS HHC, consistent with the population selected in the agency's most recent major update to existing 304(a) HHC (EPA, 2015) and the EPA's 2000 Methodology (EPA, 2000a). The EPA used the mean weight for adults ages 21 and older of 80.0 kg, based on National Health and Nutrition Examination Survey (NHANES) data from 1999 to 2006 as reported in Table 8.1 of the EPA's Exposure Factors Handbook (EPA, 2011), the EPA's most recent publication of body weight exposure factors. 4.1.2 Drinking Water Intake Rate For adults ages 21 and older, the EPA used an updated DWI of 2.3 L/d, rounded from 2.345 L/d. This DWI was estimated using the Food Commodity Intake Database consumption calculator (http://fcid.foodrisk.org8) which is based on NHANES 2005-2010 data used to develop the EPA's Exposure Factors Handbook Update to Chapter 3, Ingestion of Water and Other Select Liquids (EPA, 2019, Section 3.3.1.1). This rate represents the per capita estimate of combined direct and indirect community waterf ingestion at the 90th percentile for adults, males and females, ages 21 and older. The EPA selected the per capita rate for the updated DWI because it represents the average daily dose estimates; that is, it includes both people who drank water during the survey period and those who did not, which is appropriate for a national-scale assessment such as the development of CWA section 304(a) national human health criteria development (EPA, 2019, Section 3.2.1). The updated DWI of 2.3 L/d reflects the latest scientific knowledge in accordance with CWA 304(a)(1). The EPA's selection of the DWI of 2.3 L/d is consistent with the 2000 Methodology's selection of a rate based on per capita community water ingestion at the 86th percentile for adults e The FCID Consumption Calculator is an application that uses National Health and Nutrition Examination Survey/What We Eat in America (NHANES/WWEIA) food intake and FCID recipes to estimate food commodity consumption for the purposes of pesticide dietary exposure assessment, as well as consumption estimates for EPA's Exposure Factors Handbook (EFH) users (University of Maryland, 2024). f Community water includes direct and indirect use of tap water for household uses and excludes bottled water and other sources (EPA, 2019, Section 3.3.1.1). Direct ingestion is defined as direct consumption of water as a beverage, while indirect ingestion includes water added during food preparation (e.g., cooking, rehydration of beverages) but not water intrinsic to purchased foods (EPA, 2019, Section 3.1). 10 ------- surveyed in the U.S. Department of Agriculture's 1994-1996 Continuing Survey of Food Intake by Individuals (CSFII) analysis (EPA, 2000a, Section 4.3.2.1). 4.1.3 Fish Consumption Rate The FCR used for adults ages 21 years and older is 22.0 g/d, or 0.0220 kg/d (EPA, 2014b, Table 9a). This FCR represents the 90th percentile per capita consumption rate of fish from inland and nearshore waters for U.S. adults ages 21 years and older based on NHANES data from 2003- 2010. The 95% confidence interval (CI) of the 90th percentile per capita FCR is 19.1 g/d and 25.4 g/d, respectively. As recommended in the 2000 Methodology, the EPA used TL-specific FCRs to better represent human dietary consumption offish. An organism's trophic position in the aquatic food web can have an important effect on the magnitude of bioaccumulation of certain chemicals. The TL- specific FCRs are numbered 2, 3, and 4, and they account for different categories offish and shellfish species based on their position in the aquatic food web: TL 2 accounts for benthic filter feeders; TL 3 accounts for forage fish; and TL 4 accounts for predatory fish (EPA, 2000a). The EPA used the following TL-specific FCRs to derive the AWQC: TL 2 = 7.6 g/d (0.0076 kg/d) (95% CI [6.4, 9.1] g/d); TL 3 = 8.6 g/d (0.0086 kg/d) (95% CI [7.2, 10.2] g/d); and TL 4 = 5.1 g/d (0.0051 kg/d) (95% CI [4.0, 6.4] g/d). Each TL-specific FCR represents the 90th percentile per capita consumption rate of fish and shellfish from inland and nearshore waters from that particular TL for U.S. adults ages 21 years and older (EPA, 2014b, Tables 16a, 17a, and 18a). The sum of these three TL-specific FCRs is 21.3 g/d, which is within the 95% CI of the overall FCR of 22.0 g/d. The EPA recommends using the TL-specific FCRs when deriving AWQC; however, the overall FCR (22.0 g/d) may be used if a simplified approach is preferred. 4.2 Bioaccumulation Factor (BAF) 4.2.1 Approach Several attributes of the bioaccumulation process are important to understand when deriving national BAFs for use in developing national recommended section 304(a) AWQC. First, the term bioaccumulation refers to the uptake and retention of a chemical by an aquatic organism from all surrounding media, such as water, food, and sediment. The term bioconcentration refers to the uptake and retention of a chemical by an aquatic organism from water only. In some cases, experiments conducted in a lab that measure bioconcentration can be used to estimate the degree of bioaccumulation expected in natural conditions. However, for many chemicals, particularly those that are highly persistent and hydrophobic, the magnitude of bioaccumulation by aquatic organisms can be substantially greater than the magnitude of bioconcentration. In these cases, an assessment of bioconcentration alone underestimates the extent of accumulation in aquatic biota. Accordingly, the EPA guidelines presented in the 2000 Methodology (EPA, 2000a) emphasize using, when possible, measures of bioaccumulation as opposed to measures of bioconcentration (EPA, 2000a). The EPA estimated BAFs for this draft PFBS AWQC using the 2000 Methodology (EPA, 2000a) and the associated Technical Support Document, Volume 2: Development of National Bioaccumulation Factors (Technical Support Document, Volume 2) (EPA, 2003). Specifically, 11 ------- these documents provide a framework for identifying alternative procedures to derive national TL-specific BAFs for a chemical based on the chemical's properties (e.g., ionization and hydrophobicity), metabolism, and biomagnification potential (EPA, 2000a, 2003). As described in the 2000 Methodology, the purpose of the EPA's national BAF is to represent the long-term, average bioaccumulation potential of a chemical in aquatic organisms that are commonly consumed by humans throughout the United States (EPA, 2000a). The EPA evaluated results from field BAF and laboratory bioconcentration factor (BCF) studies on aquatic organisms commonly consumed by humans in the United States for use in developing national trophic- level BAFs. National BAFs are not intended to reflect fluctuations in bioaccumulation over short periods (e.g., a few days) because human health AWQC are generally designed to protect humans from long-term (lifetime) exposures to waterborne chemicals (EPA, 2003). The EPA followed the approach described in Figure 3-1 of the Technical Support Document, Volume 2 (EPA, 2003). The EPA used the best available data to classify each chemical according to this framework, and to derive the most appropriate BAFs following the 2000 Methodology (EPA, 2000a) and Technical Support Document, Volume 2 (EPA, 2003). Best available data consisted of peer-reviewed literature sources, government reports, and professional society proceedings, when sufficient information was provided to indicate the quality and usability of the data. The framework provides six procedures to calculate a national BAF based on the pollutant's physical and chemical properties (see Figure 1, Procedures 1-6). Each procedure contains a hierarchy of the BAF derivation methods (listed below); however, this hierarchy should not be considered inflexible (EPA, 2000a). The four methods are: 1. BAF Method. This method calculates national TL-specific BAFs using water and fish and shellfish tissue concentration data obtained from field studies. Field-measured BAFs are calculated by dividing the concentration of a contaminant in an organism by the concentration of that contaminant in the surrounding water. For nonionic organic chemicals, BAFs are normalized to allow a common basis for averaging BAFs from several studies by adjusting for the water-dissolved portions of the chemical. In order to calculate representative TL-specific national BAFs used to calculate national recommended 304(a) criteria, the EPA averaged multiple field BAFs using a geometric mean of the normalized BAFs, first by species and then by TL, to calculate the TL baseline BAFs. 2. BSAF Method. This method uses biota-sediment accumulation factors (BSAFs) to estimate bioaccumulation. While BAFs are calculated by dividing the concentration of a contaminant in an organism by the concentration of the contaminant in water, BSAFs divide the concentration in the organism by the concentration in surrounding sediments. BSAFs are useful when calculating site-specific criteria for compounds that are highly hydrophobic—these compounds have the potential to cause bioaccumulation in aquatic organisms even when concentrations in the water column are below detection limits. 12 ------- 3. BCF Method. This method estimates BAFs from laboratory-measured BCFs. Experiments designed to calculate BCFs aim to measure bioconcentration resulting from an organism's exposure to contaminated water. Unlike BAFs measured in the field, BCF experiments do not capture bioaccumulation from other routes of exposure or biomagnification (the increase in bioaccumulation at higher levels of the food chain). However, BCFs may be used to estimate bioaccumulation if a contaminant's chemical and physical properties indicate that the compound is likely to primarily accumulate in the organism via the water exposure route, and there is no evidence that the contaminant biomagnifies in the food chain. If insufficient field-collected data are available to calculate a national BAF, then the EPA may also estimate bioaccumulation using laboratory measured BCFs and a food chain multiplier term, which accounts for biomagnification. A similar process to the one described in the BAF method description (above) for normalizing of water-dissolved portions of the chemical and particulate organic carbon content is used for calculating national BAFs from laboratory-measured BCF data. Ionic organic chemicals are normalized, then multiplied by the food chain multiplier if biomagnification is likely to occur. All available BCFs are averaged using a geometric mean across species and then across TL to compute baseline BAFs. 4. Kow Method. This method predicts BAFs based on a chemical's octanol-water partition coefficient (Kow), with or without adjustment using a food chain multiplier, as described in Section 5.4 of the Technical Support Document, Volume 2 (EPA, 2003). 4.2.2 Data Selection and Evaluation The EPA conducted a systematic literature search in June 2023 of publicly available literature sources to determine whether they contained information relevant to calculating national BAFs for human health AWQC, using the 2000 Methodology and Technical Support Document, Volume 2 (EPA, 2000a, 2003). Initially, bioaccumulation data published in Burkhard (2021) was reviewed for inclusion. Burkhard (2021) evaluated bioaccumulation literature through mid-2020. To supplement this literature, a second literature search was conducted to identify additional bioaccumulation data published from 2020 through June 2023. The literature search for reporting the bioaccumulation of PFBS was implemented by developing a series of chemical-based search terms (see Appendix B) consistent with the process for derivation of BAFs used in the development of the EPA's Final Aquatic Life Criteria for perfluorooctanoic acid (PFOA) (EPA, 2024d) and PFOS (EPA, 2024e) and described in Burkhard (2021). These terms included chemical names and Chemical Abstracts Service Registry Numbers (CASRNs or CAS), synonyms, tradenames, and other relevant chemical forms (i.e., related compounds) (see Section 8). Databases searched were Current Contents, ProQuest CSA, Dissertation Abstracts, Science Direct, Agricola, Scopus, PubMed, Google Scholar, TOXNET, and UNIFY (database internal to the EPA's ECOTOX database). The supplemental literature search yielded > 10,000 results and the citation list that were further refined by excluding citations on analytical methods, human health, terrestrial organisms, bacteria, and where PFBS was not a chemical of study. The citations meeting the search criteria were reviewed for reported BAFs and/or reported concentrations in which BAFs could be calculated. Data from papers that met the inclusion and data quality screening criteria described below were extracted into the chemical dataset for PFBS. 13 ------- Specifically, studies were evaluated for inclusion in the dataset used for calculating national BAFs for PFBS using the following evaluation criteria: • Only BAF studies that included units for tissue, water, and/or BAFs were included. • Mesocosm, microcosm, and model ecosystem studies were not selected for use in calculating BAFs. • BAF studies in which concentrations in tissue and/or water were below the minimum level of detection were excluded. • Only studies performed using freshwater or brackish water were included; high salinity values were excluded. • Studies of organisms (e.g., damselfly, goby) and tissues (e.g., fish bladder) not commonly consumed by humans or not used as surrogate species for those commonly consumed by humans were excluded. Information on the ecology, physiology, and biology of the organism was used to determine whether an organism is a reasonable surrogate of a commonly consumed organisms. • Studies in which the BAFs were not found to be at steady state were excluded. • Initially, for pooled samples, averaging BAF data from multiple locations was only considered acceptable if corresponding tissue and water concentrations were available from matching locations (e.g., a BAF would not have been calculated using water and tissue samples collected from eight separate locations with tissue concentrations collected from only six of these corresponding locations). After further review, water samples averaged from samples collected between tissue sampling sites, were considered acceptable as these water samples were determined to be from the same overall spatial area of the study. In addition to the evaluation criteria listed above, PFBS bioaccumulation data were also evaluated using five study quality criteria outlined in Burkhard (2021) to further evaluate BAF literature for inclusion in the national BAF calculation (Table 1). As noted in Burkhard (2021), study quality determinations based on temporal and spatial coordination were subjective. In the absence of adequate quantifiable information regarding sample location (site coordinates for both water and tissue collection locations) or temporal coordination (specific dates of sample collection), additional follow-up with study authors was used to determine final quality values. 14 ------- Table 1. Bioaccumulation factor (BAF) study quality criteria based on suggested criteria in Burkhard (2021). Criteria 1 2 3 Number of water samples collected > 3 samples 2-3 samples 1 sample Number of organism samples collected > 3 samples 2-3 samples 1 sample Temporal coordination of water and biota samples Concurrent collection of samples Collected within a 1- year time frame Collected > 1-year time frame Spatial coordination of water and biota samples Collected from same locations Collected from reasonably close locations (1 kilometer [km]-2 km) Significantly different sampling locations General experimental design Assigned a default value of zero for studies in which tissues from individual species were identified and analyzed Assigned a value of 3 for studies in which tissues were from mixed species or reported as a taxonomic group. Notes: The scores for each BAF were totaled and used to determine the overall confidence ranking for each individual BAF. The sum of quality values for the five criteria listed in Table 1 were classified as high quality (total score of 4 or 5), medium quality (total score of 5 or 6) or low quality (total score > 7). Only high and medium quality data were included in final national BAFs calculations. 4.2.3 BAFs for PFBS Following the decision framework presented in Figure 1, the EPA selected one of the four methods to develop a national-level BAF for this chemical. Because PFBS is an organic chemical that predominantly exist in an anionic form in water (ATSDR, 2021; EPA, 2021a,b, 2024a), the BSAF and Kow methods would not be applicable. The EPA selected the BAF estimate using the BAF method (i.e., based on a field measured BAF) because sufficient field measured BAF data were available for PFBS. The national-level BAF equation adjusts the TL baseline BAFs for nonionic organic chemicals by national default values for lipid content, as well as dissolved and particulate organic carbon content. The partitioning of PFBS is related to protein binding properties (ATSDR, 2021; ECHA, 2019); therefore, the EPA did not normalize measured BAF values for PFBS using lipid content when calculating baseline and national BAFs. The EPA selected the recommended 50th percentile dissolved and particulate organic carbon content for the national-level default values which is consistent with the goal of national BAFs (i.e., as central tendency estimates), as described in Section 6.3 of the Technical Support Document, Volume 2 (EPA, 2003). Adjustment for water- dissolved portions of PFBS is applied to TL baseline BAFs (EPA, 2000a) (see Appendix B). The EPA followed the framework described in the Technical Support Document, Volume 2 (EPA, 2003), also presented in Figure 1, to select a procedure for estimating national BAFs for PFBS. 15 ------- Define Chemical of Concern I Collect & Review Data r Nonionic Organic Classify Chemical of Concern iS Hydrophobic Moderate-High T Low (Log Kow > 4) (Log Kow < 4) Ionic Organic Ionization ^— NeqliqibleZ^ i" i Yes No Inorganic & Organometallic I Procedure #1 1. Field BAF 2. BSAF 3. Lab BCF*FCM 4. Kow*FCM Procedure #3 1. Field BAF 2. Kow Procedure #5 Procedure #6 1. Field BAF or 1. Field BAF Lab BCF 2. Lab BCF*FCM Procedure #2 1. Field BAF 2. BSAF 3. Lab BCF Procedure #4 1. Field BAF or Lab BCF Figure 1. Application of the BAF framework for PFBS; gray boxes indicate steps followed based on available information for PFBS (EPA, 2000a). Based ori the characteristics for PFBS, the EPA selected Procedure 5 for deriving a national BAF value. PFBS has the following characteristics: • Ionic organic chemicals, with ionization not negligible (ATSDR, 2021; EPA, 2021a,b, 2024a). Biomagnification unlikely (Loi et al., 2011). 16 ------- The EPA was able to locate peer-reviewed, field-measured BAFs forTLs 2, 3, and 4 from the sources evaluated for which sufficient information was provided to indicate the quality and usability of the data; therefore, the EPA included only field BAF studies. The EPA used the BAF method to derive the national BAF values for PFBS: • TL 2 = 360 L/kg • TL 3 = 290 L/kg • TL 4 = 870 L/kg 5 Selection of Toxicity Value 5.1 Approach The EPA considered all available final toxicity values for both noncarcinogenic and carcinogenic toxicological effects after chronic oral exposure to develop AWQC for PFBS. As described in the 2000 Methodology (EPA, 2000a), where data are available, the EPA derives AWQC for both noncarcinogenic and carcinogenic effects and selects the more protective value for the recommended AWQC. (See Section 7, Criteria Derivation: Analysis.) For noncarcinogenic toxicological effects, the EPA uses a chronic-duration oral reference values (RfVs; RfDs or equivalent) to derive human health AWQC. An RfV is an estimate (with uncertainty spanning perhaps an order of magnitude) of a daily oral exposure of the human population to a substance that is likely to be without an appreciable risk of deleterious effects during a lifetime (EPA, 2002). An RfV may be derived from an animal toxicological study or a human epidemiological study, from which a point of departure (POD; i.e., a no-observed- adverse-effect level [NOAEL], lowest-observed-adverse-effect level [LOAEL], or benchmark dose [BMD]) can be derived. To derive the RfV, uncertainty factors are applied to the POD to reflect the limitations of the data in accordance with the EPA human health risk assessment methodology (EPA, 2014a, 2021b, 2024a). For carcinogenic toxicological effects, the EPA uses an oral CSF, when applicable and available, to derive human health AWQC. The oral CSF is an upper bound, approximating a 95% confidence limit, on the increased cancer risk from a lifetime oral exposure to a stressor. This value may also be derived from animal toxicological studies or human epidemiological studies. In developing AWQC, the EPA conducts a systematic search of peer-reviewed, publicly available final toxicity assessments to obtain the toxicity value(s) (RfV and/or CSF) for use in developing AWQC. The EPA identified toxicological assessments by systematically searching websites of the following EPA program offices, other national and international programs, and state programs in January 2024: • EPA, Office of Research and Development o Integrated Risk Information System (IRIS) program (EPA, 2024f) o Provisional Peer-Reviewed Toxicity Values (PPRTV) (EPA, 2024g) o ORD Human Health Toxicity Values (EPA, 2024h) 17 ------- • EPA, Office of Pesticide Programs (EPA, 2024i) • EPA, Office of Pollution Prevention and Toxics (EPA, 2024j) • EPA, Office of Water (EPA, 2024k) o Drinking Water Health Effects Support Documents o Toxicity Assessments • U.S. Department of Health and Human Services, Agency for Toxic Substances and Disease Registry (ATSDR, 2024) • Health Canada (HC, 2023) • California Environmental Protection Agency, Office of Environmental Health Hazard Assessment (CalEPA, 2024) After identifying and documenting all available final toxicity values, the EPA followed a systematic process to consider the identified toxicity values and select the toxicity value(s) to derive the AWQC for noncarcinogenic and carcinogenic effects. The EPA selected IRIS toxicity values to derive the draft AWQC if any of the following conditions were met: 1. The EPA's IRIS toxicological assessment was the only available source of a toxicity value. 2. The EPA's IRIS toxicological assessment was the most current source of a toxicity value. 3. The toxicity value from a more current toxicological assessment from a source other than the EPA's IRIS program was based on the same principal study and was numerically the same as an older toxicity value from the EPA IRIS program. 4. A more current toxicological assessment from a source other than the EPA's IRIS program was available, but it did not include the relevant toxicity value (chronic- duration oral RfD or CSF). 5. A more current toxicological assessment from a source other than the EPA's IRIS program was available, but it did not introduce new science (e.g., the toxicity value was not based on a newer principal study) or use a more current modeling approach compared to an older toxicological assessment from the EPA's IRIS program. The EPA selected the toxicity value from a peer-reviewed, publicly available source other than the EPA IRIS program to derive the draft AWQC if any of the following conditions were met: 1. The chemical is currently used as a pesticide, and the EPA Office of Pesticide Programs had a toxicity value that was used in pesticide registration decision-making. 2. A toxicological assessment from a source other than the EPA's IRIS program was the only available source of a toxicity value. 3. A more current toxicological assessment from a source other than the EPA's IRIS program introduced new science (e.g., the toxicity value was based on a newer principal study) or used a more current modeling approach compared to an older toxicological assessment from the EPA's IRIS program. 18 ------- 5.2 Toxicity Value for PFBS 5.2.1 Reference Dose After following the approach outlined in Section 5.1, the EPA identified the Provisional Peer- Reviewed Toxicity Values for Perfluorobutane Sulfonic Acid (PFBS) and Related Compound Potassium Perfluorobutane Sulfonate (EPA, 2021a), developed by the EPA's PPRTV program. The EPA identified a second human health assessment, Human Health Toxicity Values for Perfluorobutane Sulfonic Acid (CASRN 375-73-5) and Related Compound Potassium Perfluorobutane Sulfonate (CASRN 29420-49-3) (EPA, 2021b). These documents are identical and were identified as the most recent toxicity assessment(s) for PFBS, which use the best available science in the evaluation of noncancer risk. The EPA did not identify any other assessments that presented newer scientific information (i.e., unique RfVs) for PFBS. The EPA's final human health toxicity assessment for PFBS (EPA, 2021a,b) considered all publicly available human epidemiology, animal toxicology, and mechanistic studies that evaluated health effects after PFBS exposure. The assessment identified associations between PFBS exposure and thyroid, developmental, and kidney health effects based on toxicology studies in animals. Limited evidence from human epidemiology studies was identified; findings for thyroid or kidney health effects was equivocal, and no studies evaluating developmental effects were identified. Human epidemiology and animal toxicology studies evaluated other health effects following PFBS exposure including effects on the reproductive system, liver, and lipid and lipoprotein homeostasis, but the evidence did not support clear associations between exposure and effect (EPA, 2021a,b). The most sensitive noncancer effect observed from sufficient quality studies was an adverse developmental effect on thyroid activity, specifically decreased serum total thyroxine, in newborn mice (postnatal day [PND] 1) born to mothers that had been orally exposed to K+PFBS throughout gestation (Feng et al., 2017; EPA, 2021a,b). To develop the chronic RfD for PFBSg, the EPA derived a human equivalent dose (HED) of 0.095 mg/kg-d from BMD modeling of the critical effect in mice. The EPA then applied a composite uncertainty factor (UF) of 300 (i.e., 10x for intraspecies variability [UFh], 3x for interspecies differences [UFa], and 10x for database deficiencies [UFd] to yield the chronic oral RfD of 3 x 10"4 mg/kg-d (EPA, 2021a,b). 5.2.2 Cancer Slope Factor Under the 2005 EPA Guidelines for Carcinogen Risk Assessment (EPA, 2005), the PFBS toxicity assessment determined that there is Inadequate Information to Assess Carcinogenic Potential for PFBS (EPA, 2021a,b). Therefore, these most recent assessment did not derive a CSF for PFBS (EPA, 2021a,b). g Data for K+PFBS were used to derive the chronic RfD for the free acid (PFBS), resulting in the same value (3 x 10"4), after adjusting for differences in molecular weight between K+PFBS (338.19) and PFBS (300.10) (EPA, 2021a,b). 19 ------- 6 Relative Source Contribution (RSC) Derivation 6.1 Approach The EPA applies an RSC to the RfD when calculating an AWQC based on noncancer effects or for carcinogens that are known to act through a nonlinear mode of action to account for the fraction of an individual's total exposure allocated to AWQC-related sources (EPA, 2000a). The purpose of the RSC is to ensure that the level of a chemical allowed by a criterion (e.g., the AWQC), when combined with other identified sources of exposure (e.g., diet excluding freshwater and estuarine fish and shellfish, ambient and indoor air) common to the population of concern, will not result in exposures that exceed the RfD. In other words, the RSC is the portion of total daily exposure equal to the RfD that is attributed to consumption of ambient water (directly or indirectly in beverages like coffee tea or soup, as well as from transfer to dietary items prepared with ambient water) and fish and shellfish from inland and nearshore waters relative to other exposure sources; the remainder of the exposure equal to the RfD is allocated to other potential exposure sources. The EPA considers any potentially significant exposure source and route when deriving the RSC. The RSC is derived by applying the Exposure Decision Tree approach published in the EPA's 2000 Methodology (EPA, 2000a). The Exposure Decision Tree approach allows flexibility in the RfD apportionment among sources of exposure and considers several characteristics of the contaminant of interest, including the adequacy of available exposure data, levels of the contaminant in relevant sources or media of exposure, and regulatory agendas (i.e., whether there are multiple health-based criteria or regulatory standards for the contaminant). The RSC is developed to reflect the exposure to the U.S. general population or a sensitive population within the U.S. general population, depending on the available data. An RSC determination first requires "data for the chemical in question... representative of each source/medium of exposure and... relevant to the identified population(s)" (EPA, 2000a). The term "data" in this context is defined as ambient sampling measurements in the media of exposure, not internal human biomonitoring metrics. More specifically, the data must adequately characterize exposure distributions including the central tendency and high-end exposure levels for each source and 95% confidence intervals for these terms (EPA, 2000a). The EPA's approach recommends a "ceiling" RSC of 80% and a "floor" RSC of 20% to account for uncertainties including unknown sources of exposure, changes to exposure characteristics over time, and data inadequacies. The EPA's Exposure Decision Tree approach states that when there are insufficient environmental monitoring and/or exposure intake data to permit quantitative derivation of the RSC, the recommended RSC is 20%. In the case of AWQC development, this means that 20% of the exposure equal to the RfD is allocated to the consumption of ambient water and fish and shellfish from inland and nearshore waters and the remaining 80% is reserved for other potential sources, such as diet (excluding fish and shellfish from inland and nearshore waters), air, consumer products, etc. This 20% RSC can be replaced if sufficient data are available to develop a scientifically defensible alternative value. If scientific data demonstrating that sources and routes of exposure other than drinking water are not anticipated for a specific pollutant, 20 ------- the RSC can be raised as high as 80% based on the available data, allowing the remaining 20% for other potential sources (EPA, 2000a). Applying a lower RSC (e.g., 20%) is a more health protective approach to public health and results in a lower AWQC. To derive an RSC for PFBS, the EPA evaluated the exposure information identified through conducting prior systematic literature searches performed as part of the EPA's Maximum Contaminant Level Goals (MCLGs) for Three Individual Per- and Polyfluoroalkyl Substances (PFAS) and a Mixture of Four PFAS (EPA, 2024a), which included available information on all exposure sources and routes for PFBS. To identify information on PFBS exposure routes and sources to inform RSC determination, the EPA considered primary literature published between 2003-2020 that was collected by the EPA's Office of Research and Development as part of an effort to evaluate evidence for pathways of human exposure to eight PFAS, including PFBS. The full description of methods for peer-reviewed journal articles is available in the EPA's Maximum Contaminant Level Goals (MCLGs) for Three Individual Per- and Polyfluoroalkyl Substances (PFAS) and a Mixture of Four PFAS (EPA, 2024a). In order to consider more recently published information on PFBS exposure, the EPA incorporated the results of a date-unlimited gray literature search that was conducted in February 2022 and 2024 as well as an ad hoc process to identify relevant and more recently published peer-reviewed scientific literature. The literature resulting from the search and screening process included only final (not draft) documents and articles that were then reviewed to inform the PFBS RSC. The following description in Section 6.2 is a summary of the information provided in the Appendix of the final MCLG document three individual PFAS, including PFBS (EPA, 2024a). 6.2 Summary of Potential Exposure Sources of PFBS Other Than Water and Freshwater and Estuarine Fish/Shellfish 6.2.1 Dietary Sources PFBS was included in a suite of individual PFAS selected as part of PFAS-targeted reexaminations of samples collected for the U.S. Food and Drug Administration's Total Diet Study (FDA, 2020a,b, 2021a,b; EPA, 2024a); however, it was not detected in any of the food samples tested. It should be noted that the FDA indicated that the sample sizes were limited and that the results should not be used to draw definitive conclusions about PFAS levels or presence in the general food supply (FDA, 2023). PFBS was detected in cow milk samples collected from a farm with groundwater known to be contaminated with PFAS, as well as in produce (collard greens) collected from an area near a PFAS production plant, in FDA studies of the potential exposure of the U.S. population to PFAS (FDA 2018, 2021c). Maximum residue levels for PFBS were not found in the Global MRL Database (Bryant Christie, Inc., 2024). In addition to efforts by the FDA, 34 peer-reviewed studies conducted in North America (n = 7), Europe (n = 26), and across multiple continents (n = 1) analyzed PFBS in food items obtained from home, recreational, or commercial sources (see Table C-l in the Appendix). Food types evaluated include fruits and vegetables, grains, meat, seafood, dairy, and fats/other (e.g., eggs, spices, and oils), with seafood showing the highest levels of PFBS detected. PFBS was not detected in any of the eight studies that analyzed human milk for PFAS (not shown in Table C-l)— one in the United States (von Ehrenstein et al., 2009) and seven in Europe (Abdallah et al., 2020; 21 ------- Beser et al., 2019; Cariou et al., 2015; Karrman et al., 2007, 2010; Lankova et al., 2013; Nyberg et al., 2018). Some PFBS dietary studies use the term "seafood" to indicate fish and shellfish from ocean, freshwater, or estuarine water bodies. Information about the water bodies assessed in individual studies is reported in the articles. Of the studies conducted in North America, four U.S. studies (Blaine et al., 2014; Byrne et al., 2017; Schecter et al., 2010; Scher et al., 2018) found PFBS in at least one food item. Locations and food sources varied in these studies. In Schecter et al. (2010), PFBS was detected in cod samples but not in any of the other foods collected from Texas grocery stores. Scher et al. (2018) detected PFBS in plant parts (leaf and stem samples) analyzed from garden produce collected at homes in Minnesota within a groundwater contamination area (GCA) impacted by a former 3M PFAS production facility (PFBS concentrations ranged from ND to 0.065 ng/g). The authors suggested that the PFBS detections in plant parts were likely associated with PFAS present in irrigation water that had accumulated in produce. Blaine et al. (2014) found PFBS in radish, celery, tomato, and peas that were grown in soil amended with industrially impacted biosolids. They also found PFBS in these crops grown in soil that had received municipal biosolid applications over 20 years. In unamended control soil samples, PFBS was only detected in radish root with an average value of 22.36 ng/g (Blaine et al., 2014). In a similar study conducted by Blaine et al. (2013), PFBS was found in lettuce, tomato, and corn grown in industrially impacted biosolids-amended soils in greenhouses. Young et al. (2012) analyzed 61 raw and retail milk samples from 17 states for PFAS, but PFBS was not detected. Several peer-reviewed publications that examined PFBS concentrations in marine fish and shellfish are also available. Schecter et al. (2010) detected PFBS in cod samples, averaging 0.12 ng/g ww. In two additional studies from North America, PFBS was not detected in samples of farmed and wild-caught seafood (Chiesa et al., 2019; Young et al., 2013). Vassiliadou et al. (2015) detected PFBS in raw shrimp (from Greek markets) but did not detect PFBS in either fried shrimp, raw hake (from Greek fishing sites), or fried hake. The European Food Safety Authority reported the presence of PFBS in various food and drink items, including fruits, vegetables, cheese, and bottled water (EFSA, 2012). For average adult consumers, the estimated exposure ranges for PFBS were 0.03-1.89 ng/kg bw-d (minimum) to 0.10-3.72 ng/kg bw-d (maximum) (EFSA, 2012). Of 27 studies conducted in Europe, 12 found PFBS in at least one food type (Table C-l). Eight of the 12 studies included food samples obtained solely from markets (D'Hollander et al., 2015; Domingo et al., 2012; Eschauzier et al., 2013; Hlouskova et al., 2013; Perez et al., 2014; Scordo et al., 2020; Surma et al., 2017; Sznajder-Katarzyriska et al., 2019). Across studies, PFBS detections were found in marine fish and shellfish; other animal products such as meat, dairy, and eggs; fruits and vegetables; tap water-based beverages such as coffee; sweets; and spices. Papadopoulou et al. (2017) analyzed duplicate diet samples with PFBS detected in only one solid food sample (ND-0.001 ng/g; DF 2%; food category unspecified). Eriksson et al. (2013) evaluated foods that were farmed or freshly caught in the Faroe Islands, and only detected PFBS in cow milk (0.019 ng/g ww) and packaged dairy milk (0.017 ng/g ww) samples among the 22 ------- products analyzed. In eight of the European studies where PFBS was not detected, foods were primarily obtained from commercial sources, but wild-caught seafood was also included. In summary, in Europe and North America, PFBS has been detected in multiple food types, including fruits, vegetables, meats, marine fish and shellfish, and other fats. Although several U.S. studies have evaluated PFBS in meats, fats/oils, fruits, vegetables, and other non-seafood food types, many of these sampling efforts were localized to specific cities or markets and/or used relatively small sample sizes. Broader-scale sampling efforts will be helpful in determining the general levels of PFBS contamination in these food types, as well as the impact of known PFAS contamination sources on PFBS concentrations in foods. 6.2.2 Food Contact Materials PFBS is not authorized for use in food packaging in the United States; however, PFBS has been detected in food packaging materials in the few available studies that investigate this potential route of exposure (ATSDR, 2021; EPA, 2021a,b). In one report from the United States, PFBS was detected in fast-food packaging (7/20 samples) although the concentrations detected were not reported (Schaider et al., 2017). The EPA identified five peer-reviewed studies in Europe (conducted in Poland, Norway, Greece, Czech Republic, and Germany) which analyzed the occurrence of PFBS in food packaging or food contact materials (FCMs), such as baking papers and fast-food boxes and wrappers. Surma et al. (2015) measured levels of 10 perfluorinated compounds in three different brands of common FCMs commercially available in Poland, including wrapping papers (n = 3), breakfast bags (n = 3), baking papers (n = 3), and roasting bags (n = 3). PFBS was detected in one brand of baking paper at 0.02 picograms per square centimeter (pg/cm2), but PFBS was not detected at or below the limit of quantitation in all other FCMs. Vestergren et al. (2015) analyzed paper plates (n = 2), paper cups (n = 1), baking covers (n = 1), and baking molds (n = 1) purchased from retail stores in Troms0 and Trondheim, Norway. PFBS was detected in one paper plate at 6.9 pg/cm2. The remaining three studies did not detect PFBS in FCMs. Zafeiraki et al. (2014) analyzed FCMs made of paper, paperboard, or aluminum foil collected from a Greek market. PFBS was not detected in any of the samples of beverage cups (n = 8), ice cream cups (n = 1), fast-food paper boxes (n = 8), fast-food wrappers (n = 6), paper materials for baking (n = 2), microwave bags (n = 3), or aluminum foil bags/wrappers (n = 14). Vavrous et al. (2016) analyzed 15 samples of paper FCMs acquired from a market in the Czech Republic. FCMs included paper packages of wheat flour (n = 2), paper bags for bakery products (n = 2), sheets of paper for food packaging in food stores (n = 2), cardboard boxes for packaging of various foodstuffs (n = 3), coated bakery release papers for oven baking at temperatures up to 220 °C (n = 3), and paper filters for coffee preparation (n = 3). PFBS was not detected in any samples. Kotthoff et al. (2015) analyzed 82 samples for perfluoroalkane sulfonate (PFSA) and perfluoroalkyl carboxylic acid (PFCA) compounds in 10 consumer products including individual paper-based FCMs (n = 33) from local retailers in Germany in 2010. PFBS was not detected in paper-based FCMs. 23 ------- Overall, the few available studies conducted in the United States and Europe indicate that PFBS may be present in food packaging materials; however, further research is needed to understand which packaging materials generally contain PFBS at the highest concentrations and with the greatest frequency. There are also uncertainties related to data gaps on topics that may influence whether food packaging is a significant PFBS exposure source in humans, including differences in transfer efficiency from different packaging types directly to humans or indirectly through foodstuffs. 6.2.3 Consumer Product Uses Several studies examined a range of consumer products and found multiple PFAS, including PFBS, at various levels (Becanova et al., 2016; Favreau et al., 2016; Gremmel et al., 2016; Kotthoff et al., 2015; Liu et al., 2014; Schultes et al., 2018; van der Veen et al., 2020; Vestergren et al., 2015; Zheng et al., 2020). Two of the studies collected consumer products in the United States, five purchased consumer products in Europe, and two studies did not report the purchase location(s) of the consumer products that were tested. Zheng et al. (2020) determined the occurrence of ionic and neutral PFAS in items collected from childcare environments in the United States. Nap mats (n = 26; 20 polyurethane foam, 6 vinyl cover samples) were collected from seven Seattle childcare centers. PFBS was detected in 5% of nap mat samples at a maximum concentration of 0.04 ng/g. Liu et al. (2014) analyzed the occurrence of PFAS in commonly used consumer products (carpet, commercial carpet-care liquids, household carpet/fabric-care liquids, treated apparel, treated home textiles, treated nonwoven medical garments, floor waxes, membranes for apparel, and thread-sealant tapes) purchased from retail outlets in the United States. PFBS was detected in 100% of commercial carpet/fabric-care liquids samples (n = 2) at concentrations of 45.8 ng/g and 89.6 ng/g, in 75% of household carpet/fabric-care liquids and foams samples (n = 4) at concentrations up to 911 ng/g, in one treated apparel samples (n = 2) at a concentration of 2 ng/g, in the single treated floor wax and stone/wood sealant sample (143 ng/g, n = 2), and in the single apparel membrane sample (30.7 ng/g, n = 2). PFBS was not detected in treated home textile and upholstery (n = 2) or thread-sealant tapes and pastes (n = 2). van der Veen et al. (2020) examined the effects of weathering on PFAS content in durable water-repellent clothing collected from six suppliers in Sweden (one pair of outdoor trousers, seven jackets, four fabrics for outdoor clothes, and one pair of outdoor overalls). Two pieces of each of the 13 fabrics were cut. One piece of each fabric was exposed to elevated ultraviolet radiation, humidity, and temperature in an aging device for 300 hours (assumed lifespan of outdoor clothing); the other was not aged. Both pieces of each fabric were analyzed for ionic PFAS (including PFBS) and volatile PFAS. In general, aging of outdoor clothing resulted in increased perfluoroalkylated acid levels of 5-fold or more. For eight of 13 fabrics, PFBS was not detected before or after aging. For three fabrics, PFBS was detected before and after aging, increasing approximately 3- to 14-fold in the aged fabric (i.e., from 43 to 140 micrograms per square meter [|-ig/m2], 45 to 350 |-ig/m2, and 9.6 to 130 |-ig/m2 respectively for the three fabrics). For the remaining two fabrics, PFBS was not detected prior to aging but was detected afterward at concentrations of 0.57 and 1.7 |-ig/m2, respectively. The authors noted 24 ------- that possible explanations for this could be weathering of precursor compounds (e.g., fluorotelomer alcohols) to PFAAs such as PFBS or increased extractability due to weathering. Kotthoff et al. (2015) analyzed 82 samples for PFSA and PFCA compounds in outdoor textiles (n = 3), gloves (n = 3), carpets (n = 6), cleaning agents (n = 6), impregnating sprays (n = 3), leather (n = 13), wood glue (n = 1), ski wax (n = 13), and awning cloth (n = 1). Individual samples were bought from local retailers or collected by coworkers of the involved institutes or local clubs in Germany. The age of the samples ranged from a few years to decades. PFBS was detected in outdoor textiles (level not provided), carpet samples (up to 26.8 |ag/m2), ski wax samples (up to 3.1 micrograms per kilogram [|-ig/kg]), leather samples (up to 120 i-ig/kg), and gloves (up to 2 |-ig/kg). Favreau et al. (2016) analyzed the occurrence of 41 PFAS in a wide variety of liquid products (n = 132 consumer products, 194 total products), including impregnating agents, lubricants, cleansers, polishes, AFFFs, and other industrial products purchased from stores and supermarkets in Switzerland. PFBS was not detected in impregnation products (n = 60), cleansers (n = 24), or polishes (n = 18). PFBS was detected in 13% of a miscellaneous category of products (n = 23) that included foam-suppressing agents for the chromium industry, paints, ski wax, inks, and tanning substances, with mean and maximum concentrations of 998 and 2,992 parts per million (ppm), respectively (median = ND). The remaining two European studies from Norway (Vestergren et al., 2015) and Sweden (Schultes et al., 2018) did not detect PFBS in the consumer products analyzed. Vestergren et al. (2015) analyzed furniture textile, carpet, and clothing samples (n = 40) purchased from retail stores in Troms0 and Trondheim, Norway, while Schultes et al. (2018) determined levels of 39 PFAS in 31 cosmetic products collected in Sweden. Both studies found measurable concentrations of at least one PFAS; however, PFBS was not detected in any of the samples. Of the two studies for which purchase location(s) were not specified, Gremmel et al. (2016) determined levels of 23 PFAS in 16 new outdoor jackets since it has been shown that outdoor jackets emit PFAS to the air as well as into water during washing. The jackets were selected based on factors such as fabric and origin of production (primarily Asia, with some origins not specified). PFBS (concentration of 0.51 |-ig/m2) was only detected in one large hardshell jacket made of 100% polyester that was polyurethane-coated and finished with Teflon® (production origin unknown). Becanova et al. (2016) analyzed 126 samples of (1) household equipment (textiles, floor coverings, electrical and electronic equipment [EEE], and plastics); (2) building materials (oriented strand board, other composite wood and wood, insulation materials, mounting and sealant foam, facade materials, polystyrene, air conditioner components); (3) car interior materials; and (4) wastes of electrical and electronic equipment (WEEE) for 15 target PFAS, including PFBS. The condition (new versus used) and production year of the samples varied; the production year ranged from 1981 to 2010. The origin(s) of production were not specified. PFBS was detected in 31/55, 9/54, 7/10, and 6/7 household equipment, building materials, car interior, and WEEE samples, respectively. The highest level was 11.4 M-g/kg found in a used 1999 screen associated with WEEE. 25 ------- In summary, in the few studies available from North America and Europe, PFBS was detected in a wide range of consumer products including clothing, household textiles and products, children's products, and commercial/industrial products. However, there is some uncertainty in these results as the number and types of products tested in each study were often limited in terms of sample size. While there is evidence indicating PFBS exposure may occur through the use of or contact with consumer products, more research is needed to understand the DF and concentrations of PFBS that occur in specific products, as well as how the concentrations of PFBS change in these products with age or weathering. 6.2.4 Indoor Dust Dust ingestion may be an important exposure source of PFAS including PFBS (ATSDR, 2021), though it should be noted that dust exposure may also occur via inhalation and dermal routes. The EPA identified several studies conducted in the United States, Canada, various countries in Europe, and across multiple continents analyzed PFBS in dust of indoor environments (primarily in homes, but also schools, childcare facilities, offices, and vehicles; see Table C-2). Most of the studies sampled dust from areas not associated with any known PFAS activity or release. PFBS concentrations in dust measured in these studies ranged from ND to 170 ng/g with three exceptions: two studies (Kato et al., 2009; Strynar and Lindstrom, 2008) reported maximum PFBS concentrations greater than 1,000 ng/g in dust from homes and daycare centers, and a third study (Huber et al., 2011) reported a PFBS concentration of 1,089 ng/g in dust from a storage room that had been used to store "highly contaminated PFC [polyfluorinated compounds] samples and technical mixtures for several years." Of the two available studies that measured PFBS in dust from vehicles, one (in the United States) detected no PFBS (Fraser et al., 2013) and the other (in Ireland) reported a DF of 75% and PFBS concentrations ranging from ND to 170 ng/g (Harrad et al., 2019). One U.S. study, Scher et al. (2019) evaluated indoor dust from 19 homes in Minnesota within a GCA impacted by the former 3M PFAS production facility. House dust samples were collected from both interior living rooms and entryways to the yard. The DFs for PFBS were 16% and 11% for living rooms and entryways, respectively, and a maximum PFBS concentration of 58 ng/g was reported for both locations. Haug et al. (2011) indicated that house dust concentrations are likely influenced by a number of factors related to the building (e.g., size, age, floor space, flooring type, ventilation); the residents or occupants (e.g., number of people, housekeeping practices, consumer habits such as buying new or used products); and the presence and use of certain products (e.g., carpeting, carpet or furniture stain-protective coatings, waterproofing sprays, cleaning agents, kitchen utensils, clothing, shoes, cosmetics, insecticides, electronic devices). In addition, the extent and use of the products affects the distribution patterns of PFAS in dust of these buildings. At this time, there is uncertainty regarding the extent of human exposure to PFBS through indoor dust compared with other exposure pathways. 26 ------- 6.2.5 Air PFAS have been released to air from WWTPs, waste incinerators, and landfills (EPA, 2016). ATSDR (2021) noted that PFAS have been detected in particulates and in the vapor phase in air and can be transported long distances via the atmosphere; they have been detected at low concentrations in areas as remote as the Arctic and ocean waters. However, the EPA's Toxic Release Inventory did not report release data for PFBS in 2020 (EPA, 20241). In addition, PFBS is not listed as a hazardous air pollutant (EPA, 2024m). 6.2.5.1 Indoor Air No studies from the U.S. reporting levels of PFBS in indoor air were identified from the peer- reviewed or gray literature. However, the EPA identified studies from Europe that are summarized below. These three studies were conducted in Norway (Barber et al., 2007), Spain (Jogsten et al., 2012), and Ireland (Harrad et al., 2019). In Norway, neutral and ionic PFAS were analyzed in four indoor air samples collected from homes in Troms0 (Barber et al., 2007). PFBS levels were below the limit of quantitation. The authors noted that measurable amounts of other ionic PFAS were found in indoor air samples, but levels were not significantly elevated above levels in outdoor air. In Spain, Jogsten et al. (2012) collected indoor air samples (n = 10) from selected homes in Catalonia and evaluated levels of 27 perfluorinated chemicals (PFCs). PFBS was not detected. In Ireland, Harrad et al. (2019) measured eight target PFAS in air from cars (n = 31), home living rooms (n = 34), offices (n = 34), and school classrooms (n = 28). PFBS was detected in all four indoor microenvironments, at DFs of 53%, 90%, 41%, and 54% in samples from homes, cars, offices, and classrooms, respectively. The mean (maximum) concentrations were 22 (270) picograms per cubic meter (pg/m3) in homes, 54 (264) pg/m3 in cars, 37 (313) pg/m3 in offices, and 36 (202) pg/m3 in classrooms. There is some evidence from European studies indicating PFBS exposure via indoor air. However, further research is needed to understand the DF and concentrations of PFBS that occur in indoor environments in the United States. 6.2.5.2 Ambient Air Similar to studies on indoor PFBS air concentrations, no studies from the U.S. reporting levels of PFBS in ambient air were identified from the peer-reviewed or gray literature. Four studies conducted across Europe (Barber et al., 2007; Beser et al., 2011; Harrad et al., 2020; Jogsten et al., 2012) and one study conducted in Canada (Ahrens et al., 2011) analyzed ambient air samples for PFBS. Two of the studies (Barber et al., 2007; Harrad et al., 2020) found detectable levels of PFBS in outdoor air. Barber et al. (2007) collected air samples from four field sites in Europe (one semirural site [Hazelrigg] and one urban site [Manchester] in the United Kingdom, one rural site from Ireland, and one rural site from Norway) for analysis of neutral and ionic PFAS. Authors did not indicate whether any of the sites had a history of local PFAS-related activities (e.g., AFFF usage, PFAS manufacturing or use). PFBS was detected in the particle phase of outdoor air samples during one of the two sampling events in Manchester at 2.2 pg/m3 and one of the two sampling events in Hazelrigg at 2.6 pg/m3. PFBS was not detected above the 27 ------- method quantification limit at the Ireland and Norway sites. Harrad et al. (2020) measured PFBS in air near 10 Irish municipal solid waste landfills located in nonindustrial areas. Air samples were collected upwind and downwind of each landfill. PFBS was detected in more than 20% of the samples, with mean concentrations (ranges) at downwind and upwind locations of 0.50 (< 0.15-1.4) pg/m3 and 0.34 (< 0.15-1.2) pg/m3, respectively. Beser et al. (2011) and Jogsten et al. (2012) did not detect PFBS in ambient air samples in Spain. Beser et al. (2011) analyzed fine airborne particulate matter (PM 2.5) in air samples collected from five stations located in Alicante province, Spain (three residential, one rural, one industrial) to determine levels of 12 ionic PFAS. PFBS was below the method quantification limit at all five locations. Jogsten et al. (2012) did not detect PFBS in ambient air samples collected outside homes in Catalonia, Spain. In the one study identified from North America, Ahrens et al. (2011) determined levels of PFAS in air around a WWTP and two landfill sites in Canada. PFBS was not detected in any sample above the MDL. PFBS has been detected in Artie air in one study, with a DF of 66% and mean concentration of 0.1 pg/m3 (Arp and Slinde, 2018; Wong et al., 2018). As with exposure to PFBS via indoor air, there is some evidence from European studies indicating PFBS is present in some ambient air samples. Further research is needed to understand the DF and concentrations of PFBS that occur in ambient environments in the United States. 6.2.6 Soil PFBS can be released into soil from manufacturing facilities, industrial uses, fire/crash training sites, and biosolids containing PFBS (ATSDR, 2021; EPA, 2021a,b). The EPA identified 16 studies that evaluated the occurrence of PFBS and other PFAS in soil, with studies conducted in the United States, Canada, and Europe (see Table C-3). Two U.S. studies and two Canadian studies (Blaine et al., 2013; Cabrerizo et al., 2018; Dreyer et al., 2012; Venkatesan and Halden, 2014) were conducted in areas not reported to be associated with any known PFAS release or were experimental studies conducted at research facilities. At these sites, PFBS levels were low (< 0.10 ng/g) or below detection limits in non-amended or control soils. Two U.S. studies by Scher et al. (2018, 2019) evaluated soils at homes in Minnesota within and outside of a GCA impacted by a former 3M PFAS production facility; for sites within the GCA, one of the studies reported a DF of 10% and a 90th percentile PFBS concentration of 0.02 ng/g, and the other reported a DF of 9% and a maximum PFBS concentration of 0.017 ng/g. For sites outside of the GCA, the DF was 17% and the maximum PFBS concentration was 0.031 ng/g. Three U.S. studies and one Canadian study analyzed soils potentially impacted by AFFF used to fight fires—one at U.S. Air Force installations with historic AFFF use (Anderson et al., 2016), two at former fire training sites (Eberle et al., 2017; Nickerson et al., 2020), and another at the site of a train derailment and fire in Canada (Mejia-Avendano et al., 2017). In these four studies, DFs ranged from 35% to 100%. PFBS concentrations in the study of the U.S. Air Force installations ranged from ND-79 ng/g, and PFBS concentrations ranged from ND-58.44 ng/g at one fire training site 28 ------- (Nickerson et al., 2020). The study of the other fire training site measured PFBS pretreatment (0.61-6.4 ng/g) and posttreatment (0.07-0.83 ng/g) (Eberle et al., 2017). The DFs and range of PFBS concentrations measured in soils at the site of the train derailment were 75% DF and ND- 3.15 ng/g, respectively, for the AFFF run-off area (measured in 2013, the year of accident) and 36% DF and ND-1.25 ng/g, respectively, at the burn site and adjacent area (measured in 2015) (Mejia-Avendano et al., 2017). Of the six European studies, one study (Harrad et al., 2020) analyzed soil samples collected upwind and downwind of 10 municipal solid waste landfills in Ireland and found PFBS levels to be higher in soils from downwind locations. Based on the overall study findings, however, the authors concluded there was no discernible impact of the landfills on concentrations of PFAS in soil surrounding these facilities. Gr0nnestad et al. (2019) investigated soils from a skiing area in Norway to elucidate exposure routes of PFAS into the environment from ski products, such as ski waxes. The authors found no significant difference in mean total PFAS in soil samples from the Granasen skiing area and the Jonsvatnet reference area but noted that the skiing area samples were dominated by long-chain PFAS (C8-C14; > 70%) and the reference area samples were dominated by short-chain PFAS (> 60%), which included PFBS. A study in Belgium (Groffen et al., 2019) evaluated soils collected at a 3M fluorochemical plant in Antwerp and at four sites located at increasing distances from the plant. PFBS levels were elevated at the plant site and decreased with increasing distance from the plant. The other three studies analyzed soil samples from areas near firefighting training sites in Norway and France and reported PFBS concentrations varying from ND to 101 ng/g dry weight (Dauchy et al., 2019; Hale et al., 2017; Skaar et al., 2019). A U.S. study of biosolid samples from 94 WWTPs across 32 states and the District of Columbia detected PFBS in 60% of samples at a mean concentration (range) of 3.4 (2.5-4.8) ng/g (Venkatesan and Halden, 2013). PFBS has been detected in drinking water wells, food types, and plant samples from soils or fields that have received biosolids applications that were industrially impacted (Blaine et al., 2013, 2014; Lindstrom et al., 2011). In summary, results of some available studies suggest that proximity to a PFAS production facility or a site with historical AFFF use or firefighting is correlated with increased PFBS soil concentrations compared to soil from sites not known to be impacted by PFAS. However, few available studies examined PFBS concentrations in soils not known to have nearby sources of PFBS. Additional research is needed that quantifies ambient levels of PFBS in soils in the United States. 6.2.7 Summary and Recommended RSC for PFBS As mentioned above, the scope of exposure sources considered for the draft recommended human health AWQC is limited to surface water used for drinking water and the consumption of freshwater/estuarine fish and shellfish (EPA, 2000a), consistent with previous human health AWQC (EPA, 2015). The EPA followed the Exposure Decision Tree approach to determine the RSC for PFBS (EPA, 2000a; see Figure 2). 29 ------- Figure 2. RSC exposure decision tree framework for PFBS; figure adapted from EPA (2000a) with gray boxes indicating key decision points for this chemical. To identify the population(s) of concern (Box 1, Figure 2), the EPA first identified potentially sensitive subpopulations or life stages based on the PFBS exposure interval in the critical study from which the critical effect (adverse developmental effect on thyroid activity) was selected for RfD derivation in the PFBS toxicity assessment (EPA, 2021a,b). Since the critical effect is the most sensitive adverse health effect that was identified from the available data of sufficient quality, then the exposure interval may be a sensitive window of exposure. The exposure interval of the critical study in rodents corresponds to the following two potentially sensitive human life stages; women of childbearing age who may be or become pregnant; and pregnant women and their developing fetus. However, limited information was available regarding specific PFBS exposure in these two life stages from different environmental sources. Therefore, the EPA considered exposures in the general U.S. population, ages 21 years of age and older, which includes these two life stages. Second, the EPA identified PFBS relevant exposure sources/pathways (Box 2, Figure 2), including nonfish (except marine) dietary consumption, incidental oral consumption via dust, consumer products, and soil or dermal exposure via soil, consumer products, and dust, and 30 ------- inhalation exposure via indoor or ambient air. Several of these may be potentially significant exposure sources. Third, the EPA evaluated whether adequate data were available to describe the central tendencies and high-end exposures for all potentially significant exposure sources and pathways (Box 3, Figure 2). The EPA determined that there were inadequate quantitative data to describe the central tendencies and high-end estimates for all of the potentially significant sources. For example, studies from Canada and Europe indicate that indoor and ambient air may be a significant source of exposure to PFBS. At the time of the literature search, the EPA was unable to identify studies assessing PFBS concentrations in indoor or ambient air samples from the United States and therefore, the agency does not have adequate quantitative data to describe the central tendency and high-end estimate of exposure for this potentially significant source in the U.S. population. Fourth, the agency determined whether there were sufficient data, physical/chemical property information, fate and transport information, and/or generalized information available to characterize the likelihood of exposure to relevant sources (Box 4, Figure 2). Sufficient information on PFBS was available to characterize the likelihood of exposure. The agency relied on the studies summarized above to determine if there are potential uses/sources of PFBS other than AWQC-related sources (Box 6, Figure 2). There are significant known or potential uses/sources of PFBS other than AWQC-related sources. Based on this information, the next step was to determine if adequate information was available on PFBS to characterize each source/pathway of exposure (Box 8a, Figure 2). The EPA determined there is not enough information available on each source to make a quantitative characterization of exposure among exposure sources. For example, there are several studies from the U.S. indicating that PFBS may occur in dust sampled from various microenvironments (e.g., homes, offices, daycare centers, vehicles). However, the majority of studies sampled in only one location and few studies examined dust samples outside of the home (e.g., one study from the U.S. assessed PFBS occurrence in dust sampled from vehicles). Additionally, though several studies from around the U.S. measured PFBS concentrations in dust from houses, the detection frequencies in these studies varied widely (from 3% to 59%) and may be a result of uncertainties including home characteristics, behaviors of the residents, and the presence or absence of PFBS- containing materials or products (Haug et al., 2011). Therefore, it is not possible to determine whether dust, food sources other than freshwater/estuarine fish and shellfish, or consumer products, may be major or minor contributors to total PFBS exposure. Therefore, the data are insufficient to allow for quantitative characterization of the different exposure sources. The EPA's Exposure Decision Tree approach states that when there is insufficient environmental and/or exposure data to permit quantitative derivation of the RSC, the recommended RSC for the general population is 20% (EPA, 2000a). Thus, the EPA recommends an RSC of 20% (0.20) for PFBS for AWQC for both the water plus organism AWQC as well as the organism only AWQC (Box 8b, Figure 2). 31 ------- 7 Criteria Derivation: Analysis Table 2 summarizes the input parameters used to derive the draft recommended human health AWQC that are protective of exposure to PFBS from consuming drinking water and/or eating Table 2. Input parameters for the human health AWQC for PFBS. Input Parameter Value RfD 0.0003 mg/kg-d CSF No data RSC 0.20 BW 80.0 kg DWI 2.3 L/d FCR TL 2 0.0076 kg/d TL 3 0.0086 kg/d TL 4 0.0051 kg/d BAF TL 2 360 L/kg TL 3 290 L/kg TL 4 870 L/kg Notes: RfD = reference dose; CSF = cancer slope factor; RSC = relative source contribution; BW = bodyweight; DWI = drinking water intake; FCR = fish consumption rate; TL = trophic level; BAF = bioaccumulation factor. fish and shellfish (organisms) from inland and nearshore waters. The criteria calculations are presented below. These criteria recommendations are based on the 2000 Methodology (EPA, 2000a) and the toxicity and exposure assumptions described above (see Section 4, AWQC Input Parameters; Section 5, Selection of Toxicity Value; and Section 6, Relative Source Contribution Derivation). 7.1 AWQC for Noncarcinogenic Toxicological Effects For consumption of water and organisms: AWQC (|-ig/L) = RfD (mg/kg-d) x RSC x BW (kg) x 1.000 (ug/mg) DWI (L/d) + £?=2 (FCRi (kg/d) x BAFi (L/kg)) = 0.0003 mg/kg-d x 0.20 x 80.0 kg x 1,000 |-ig/mg 2.3 L/d + ((0.0076 kg/d x 360 L/kg) + 0.0086 kg/d x 290 L/kg) + (0.0051 kg/d x 870 L/kg)) = 0.4011 ng/L = 0.4 |ag/L (rounded) 32 ------- For consumption of organisms only: AWQC (|-ig/L) = RfD (mg/kg-d) x RSC x BW (kg) x 1.000 (ug/mg) Sf=2 (FCRi (kg/d) x BAFi (L/kg)) = 0.0003 mg/kg-d x 0.20 x 80.0 kg x 1,000 |ag/mg (0.0076 kg/d x 360 L/kg) + (0.0086 kg/d x 290 L/kg) + (0.0051 kg/d x 870 L/kg) = 0.4965 |ag/L = 0.5 |ag/L (rounded) 7.2 AWQC for Carcinogenic Toxicological Effects The PFBS toxicity assessments determined that there is Inadequate Information to Assess Carcinogenic Potential for PFBS (EPA, 2021a,b; see Section 5, Selection of Toxicity Value). The EPA derives cancer-based HHC for contaminants that have been determined to be Carcinogenic to Humans or Likely to Be Carcinogenic to Humans (EPA, 2000a,c). Therefore, the EPA did not derive AWQC for carcinogenic toxicological effects. 7.3 AWQC Summary for PFBS The EPA derived the draft recommended AWQC for PFBS using a noncarcinogenic toxicity endpoint. The human health AWQC for noncarcinogenic effects for PFBS are 0.4 |ig/L (400 ng/L) for consumption of water and organisms and 0.5 |ig/L (500 ng/L) for consumption of organisms only (Table 3). The EPA evaluated the use of exposure factors relevant to sensitive subpopulations based on the critical effect(s) used to derive the RfD (Appendix D). Based on the results of this evaluation, the criteria based on exposure factors for the general adult (> 21 years of age) population are the most health protective. Under the EPA's recently finalized Method 1633 (EPA, 2024n) for aqueous samples, the level of quantification (LOQ) representing the observed LOQs in the multi-laboratory validation study, range from 1 to 4 ng/L for PFBS. The pooled MDL for PFBS is 0.37 ng/L. The pooled MDL value is derived from the multi-laboratory validation study using MDL data from eight laboratories and represents the sensitivity that should be achievable in a well-prepared laboratory but may not represent the actual MDL used for data reporting or data quality assessments (EPA, 2024n). The MDLs and ranges presented here provide a reference for comparison of analytical concentrations and recommended criteria. Table 3. Summary of the EPA's recommended human health AWQC for PFBS chemicals. Human Health AWQC Water and Organism 0.4 ng/L (400 ng/L) Organism Only 0.5 ng/L (500 ng/L) 33 ------- 8 Consideration of Noncancer Health Risks from PFAS Mixtures The EPA recently released its final Framework for Estimating Noncancer Health Risks Associated with Mixtures of Per- and Polyfluoroalkyl Substances (PFAS) (referred to here as the PFAS mixtures framework; EPA, 2024o). The PFAS mixtures framework describes three flexible, data- driven approaches that facilitate practical component-based mixtures evaluation of two or more PFAS based on dose additivity, consistent with the EPA's Guidelines for the Health Risk Assessment of Chemical Mixtures (EPA, 1986) and Supplementary Guidance for Conducting Health Risk Assessment of Chemical Mixtures (EPA, 2000d). The approaches described in the EPA PFAS mixtures framework may support interested federal, state, and Tribal partners, as well as public health experts and other stakeholders to assess the potential noncancer human health hazards and risks associated with PFAS mixtures. The EPA is providing an illustration of one approach which could be applied to PFAS mixture HHC derivation. The PFAS mixtures framework underwent peer review by the EPA Science Advisory Board (EPA, 2022b) and public review and the EPA responded to comments (EPA, 2024p). The public comment period ended on May 30, 2023. The public docket can be accessed at www.regulations.gov under Docket ID: EPA-HQ-OW-2022-0114. Dose additivity means that the combined effect of the component chemicals in a mixture is equal to the sum of the individual doses or concentrations scaled for potency. As noted in the PFAS mixtures framework, exposure to a number of individual PFAS has been shown to elicit the same or similar profiles of adverse effects in various organs and systems. Many toxicological studies of PFAS as well as other classes of chemicals support the health-protective conclusion that chemicals that elicit the same or similar observed adverse effects following individual exposure should be assumed to act in a dose-additive manner when in a mixture unless data demonstrate otherwise (EPA, 2024o). Importantly, few studies have examined the toxicity of PFAS mixtures, particularly with component chemical membership and proportions that are representative of the diverse PFAS mixtures that occur in the environment. Mixtures assessments for chemicals that share similar adverse health effects, and therefore assume dose additivity, typically apply component-based assessment approaches. The Hazard Index (HI) approach is one of the component-based mixtures assessment approaches described in the PFAS mixtures framework. In order to support states and Tribes interested in addressing potential noncancer risks of PFAS mixtures, the application of the HI approach for deriving HHC for mixtures is described below. States and authorized Tribes may choose to adopt this approach to derive HHC for PFAS mixtures. Use of the HI approach to assess risks associated with PFAS mixtures was supported by the EPA Science Advisory Board (EPA, 2022b). In the HI approach (see PFAS mixtures framework; EPA, 2024o), a hazard quotient (HQ) is calculated as the ratio of human exposure (E) to a human health-based toxicity value (e.g., reference value [RfV]) for each mixture component chemical (i) (EPA, 1986). The HQs for the component chemicals are then summed to derive a mixture-specific HI (for the specified exposure route/medium). Since the HI is unitless, the E and the RfV inputs to the HI formula must be expressed in the same dose units (e.g., mg/L) (Eq. 5). For example, in the context of the 34 ------- human health criteria, HQs for each individual PFAS are calculated by dividing the measured ambient water concentration of each component PFAS (e.g., expressed as |ag/L) by its corresponding human health criterion (e.g., expressed as |ag/L), and the resulting component PFAS HQs are summed to yield the PFAS mixture HI (Eqs. 5-7). Either water-plus-organism or organism-only HHC can be used in the PFAS mixtures HI approach; however, the type of HHC selected for HI calculation should be consistent. Because cancer data are lacking for most PFAS, the HI approach is currently recommended for PFAS HHC based on noncancer effects. A hypothetical example is included below to illustrate using the HI approach to derive an HHC for a mixture of three PFAS. A PFAS mixture HI exceeding 1 indicates that co-occurrence of two or more PFAS in a mixture in ambient water exceeds the health-protective level(s), indicating potential health risks. Some individual PFAS have HHC below the analytical MDLs (e.g., PFOA, PFOS). If one such PFAS is included as a component PFAS in the HI approach, then any detectable level of that component PFAS in surface water will result in a component HQ greater than 1, and thus, an HI greater than 1 for the PFAS mixture. HI = Zr=iHQi =ZF=1^ (Eq.5) HI = HQpfbs + HQPFASx + HQPFASy (Eq. 6) j_jj /[PFBSambjent water A _|_ / PFASX ambient water N _|_ /[PFASy^ambient water]\ y\ V [PFBShhc] / V [PFASx,hhc] / V [PFASy hhc] / Where: HI = hazard index n = the number of component (i) PFAS HQi = hazard quotient for component (i) PFAS Ei = human exposure for component (i) PFAS HHQ = human health criterion for component PFAS (i) HQpfas = hazard quotient for a given individual PFAS PFASx= Hypothetical PFAS PFASy= Hypothetical PFAS [PFASambient water] = concentration of a given PFAS in ambient water [PFAShhc] = water-plus-organism HHC or organism-only HHC for a given PFAS 9 Chemical Name and Synonyms Perfluorobutane Sulfonic Acid (CASRN 375-73-5) Potassium Perfluorobutane Sulfonate (CASRN 29420-49-3) PFBS K+PFBS 1,1,2,2,3,3,4,4,4-Nonafluoro-l-butanesulfonic acid 1-Perfluorobutanesulfonic acid 35 ------- Nonafluoro-l-butanesulfonic acid • Nonafluorobutanesulfonic acid • Perfluorobutanesulfonic acid • 1,1,2,2,3,3,4,4,4-Nonafluorobutane-l-sulphonic acid • Perfluorobutanesulfonate • Perfluorobutane sulfonate • 1-Butanesulfonic acid, 1,1,2,2,3,3,4,4,4-nonafluoro- • 1-Butanesulfonic acid, nonafluoro- • Perfluoro-l-butanesulfonate • Perfluorobutylsulfonate 10 References 3M (3M Company). 2002. Environmental, Health, Safety, and Regulatory (EHSR) Profile of Perfluorobutane Sulfonate (PFBS). Accessed February 2024. http://multimedia.3m.com/mws/media/172303O/ehsr-profile-of- perfluorobutanesulfonate-pfbs.pdf. Abdallah, M.A., N. Wemken, D.S. Drage, C. Tlustos, C. Cellarius, K. Cleere, J.J. Morrison, S. Daly, M.A. Coggins, and S. Harrad. 2020. Concentrations of perfluoroalkyl substances in human milk from Ireland: Implications for adult and nursing infant exposure. Chemosphere 246:125724. http://dx.doi.Org/10.1016/i.chemosphere.2019.125724. AECOM. 2019. Perfluorobutane Sulfonic Acid (PFBS) Chemistry, Production, Uses, and Environmental Fate in Michigan. Project Number 60560354. Accessed April 2024. https://www. michigan.gov/-/med ia/Project/Websites/PFAS- Response/Workgroups/Groundwater/Report-2019-09-23-PFBS-Chemistry-Production- Uses-Fate.pdf?rev=c8b3fl7030a0459ca02366df9c712648. Ahrens, L., S. Felizeter, R. Sturm, Z. Xie, and R. Ebinghaus. 2009a. Polyfluorinated compounds in waste water treatment plant effluents and surface waters along the River Elbe, Germany. Marine Pollution Bulletin 58:1326-1333. http://dx.doi.Org/10.1016/i.marpolbul.2009.04.028. Ahrens, L., M. Plassmann, Z. Xie, and R. Ebinghaus. 2009b. Determination of polyfluoroalkyl compounds in water and suspended particulate matter in the river Elbe and North Sea, Germany. Frontiers of Environmental Science & Engineering in China 3:152-170. http://dx.doi.org/10.1007/sll783-009-0Q21-8. 36 ------- Ahrens, L., M. Shoeib, T. Harner, S. Lee, R. Guo, and E.J. Reiner. 2011. Wastewater treatment plant and landfills as sources of polyfluoroalkyl compounds to the atmosphere. Environmental Science & Technology 45:8098-8105. http://dx.doi.org/10.1021/eslQ36173. Anderson, R.H., G.C. Long, R.C. Porter, and J.K. Anderson. 2016. Occurrence of select perfluoroalkyl substances at U.S. Air Force aqueous film-forming foam release sites other than fire-training areas: Field-validation of critical fate and transport properties. Chemosphere 150:678-685. http://dx.doi.Org/10.1016/i.chemosphere.2016.01.014. Appleman, T.D., C.P. Higgins, O. Quinones, B.J. Vanderford, C. Kolstad, J.C. Zeigler-Holady, and E.R. Dickenson. 2014. Treatment of poly- and perfluoroalkyl substances in U.S. full-scale water treatment systems. Water Research 51:246-255. http://dx.doi.Org/10.1016/i.watres.2013.10.067. Arp, H.P., and G.A. Slinde. 2018. PFBS in the Environment: Monitoring and Physical-Chemical Data Related to the Environmental Distribution of Perfluorobutanesulfonic Acid. Report M-1122. Norwegian Environmental Agency. Accessed April 2024. https://www.miliodirektoratet.no/globalassets/publikasioner/M1122/M1122.pdf. ATSDR (Agency for Toxic Substances and Disease Registry). 2021. Toxicological Profile for Perfluoroalkyls. U.S. Department of Health and Human Services, ATSDR, Atlanta, GA. Accessed January 2024. https://www.atsdr.cdc.gov/toxprofiles/tp200.pdf. ATSDR (Agency for Toxic Substances and Disease Registry). 2024. Toxicological Profiles. U.S. Department of Health and Human Services, ATSDR, Atlanta, GA. Accessed January 2024. https://www.atsdr.cdc.gov/toxicological-profiles/about/index.html. Bach, C., X. Dauchy, V. Boiteux, A. Colin, J. Hemard, V. Sagres, C. Rosin, and J.F. Munoz. 2017. The impact of two fluoropolymer manufacturing facilities on downstream contamination of a river and drinking water resources with per- and polyfluoroalkyl substances. Environmental Science and Pollution Research 24:4916-4925. http://dx.doi.org/10.1007/sll356-016-8243-3. Barber, J.L., U. Berger, C. Chaemfa, S. Huber, A. Jahnke, C. Temme, and K.C. Jones. 2007. Analysis of per- and polyfluorinated alkyl substances in air samples from Northwest Europe. Journal of Environmental Monitoring 9:530-541. http://dx.doi.org/10.1039/b7Q1417a. Barreca, S., M. Busetto, L. Colzani, L. Clerici, V. Marchesi, L. Tremolada, D. Daverio, and P. Dellavedova. 2020. Hyphenated high performance liquid chromatography-tandem mass spectrometry techniques for the determination of perfluorinated alkylated substances in Lombardia region in Italy, profile levels and assessment: One year of monitoring activities during 2018. Separations 7(1):17. http://dx.doi.org/10.3390/separations7010Q17. 37 ------- Becanova, J., L. Melymuk, S. Vojta, K. Komprdova, and J. Klanova. 2016. Screening for perfluoroalkyl acids in consumer products, building materials and wastes. Chemosphere 164:322-329. http://dx.doi.Org/10.1016/i.chemosphere.2016.08.112. Beser, M.I., 0. Pardo, J. Beltran, and V. Yusa. 2011. Determination of per- and polyfluorinated substances in airborne particulate matter by microwave-assisted extraction and liquid chromatography-tandem mass spectrometry. Journal of Chromatography A 1218:4847- 4855. http://dx.doi.Org/10.1016/i.chroma.2011.04.082. Beser, M.I., O. Pardo, J. Beltran, and V. Yusa. 2019. Determination of 21 perfluoroalkyl substances and organophosphorus compounds in breast milk by liquid chromatography coupled to orbitrap high-resolution mass spectrometry. Analytical Chimica Acta 1049:123-132. http://dx.doi.Org/10.1016/i.aca.2018.10.033. Blaine, A.C., C.D. Rich, L.S. Hundal, C. Lau, M.A. Mills, K.M. Harris, and C.P. Higgins. 2013. Uptake of perfluoroalkyl acids into edible crops via land applied biosolids: Field and greenhouse studies. Environmental Science & Technology 47:14062-14069. http://dx.doi.org/10.1021/es403094q. Blaine, A.C., C.D. Rich, E.M. Sedlacko, L.S. Hundal, K. Kumar, C. Lau, M.A. Mills, K.M. Harris, and C.P. Higgins. 2014. Perfluoroalkyl acid distribution in various plant compartments of edible crops grown in biosolids-amended soils. Environmental Science & Technology 48:7858-7865. http://dx.doi.org/10.1021/es500Q16s. Boiteux, V., X. Dauchy, C. Rosin, and J.F. Munoz. 2012. National screening study on 10 perfluorinated compounds in raw and treated tap water in France. Archives of Environmental Contamination and Toxicology 63:1-12. http://dx.doi.org/10.1007/sQ0244-012-9754-7. Boiteux, V., X. Dauchy, C. Bach, A. Colin, J. Hemard, V. Sagres, C. Rosin, and J.F. Munoz. 2017. Concentrations and patterns of perfluoroalkyl and polyfluoroalkyl substances in a river and three drinking water treatment plants near and far from a major production source. Science of the Total Environment 583:393-400. http://dx.doi.Org/10.1016/i.scitotenv.2017.01.079. Bryant Christie, Inc. 2024. Global MRL Database: Pesticide MRLs Page. Bryant Christie, Inc., Seattle, WA. Accessed January 2024. https://bcglobal.bryantchristie.eom/d b#/pesticides/query. Burkhard, L.P. 2021. Evaluation of published bioconcentration factor (BCF) and bioaccumulation factor (BAF) data for per- and polyfluoroalkyl substances across aquatic species. Environmental Toxicology and Chemistry 40(6):1530-1543. https://setac.onlinelibrarv.wiley.com/doi/10.1002/etc.5010. 38 ------- Byrne, S., S. Seguinot-Medina, P. Miller, V. Waghiyi, F.A. von Hippel, C.L. Buck, and D.O. Carpenter. 2017. Exposure to polybrominated diphenyl ethers and perfluoroalkyl substances in a remote population of Alaska Natives. Environmental Pollution 231:387- 395. http://dx.doi.Org/10.1016/i.envpol.2017.08.020. Cabrerizo, A., D.C.G. Muir, A.O. De Silva, X. Wang, S.F. Lamoureux, and M.J. Lafreniere. 2018. Legacy and emerging persistent organic pollutants (POPs) in terrestrial compartments in the high Arctic: Sorption and secondary sources. Environmental Science & Technology 52:14187-14197. http://dx.doi.org/10.1021/acs.est.8b05011. CalEPA (California Environmental Protection Agency). 2024. Public Health Goals (PHGs). CalEPA, Office of Environmental Health Hazard Assessment. Accessed January 2024. https://oehha.ca.gov/water/public-health-goals-phgs. Cariou, R., B. Veyrand, A. Yamada, A. Berrebi, D. Zalko, S. Durand, C. Pollono, P. Marchand, J.C. Leblanc, J.P. Antignac, and B. Le Bizec. 2015. Perfluoroalkyl acid (PFAA) levels and profiles in breast milk, maternal and cord serum of French women and their newborns. Environment International 84:71-81. http://dx.doi.Org/10.1016/i.envint.2015.07.014. Chiesa, L.M., M. Nobile, F. Ceriani, R. Malandra, F. Arioli, and S. Panseri. 2019. Risk characterisation from the presence of environmental contaminants and antibiotic residues in wild and farmed salmon from different FAO zones. Food Additives and Contaminants: Part A 36:152-162. http://dx.doi.org/10.1080/19440049.2Q18.1563723. Dauchy, X., V. Boiteux, C. Bach, C. Rosin, and J.F. Munoz. 2017. Per- and polyfluoroalkyl substances in firefighting foam concentrates and water samples collected near sites impacted by the use of these foams. Chemosphere 183:53-61. http://dx.doi.Org/10.1016/i.chemosphere.2017.05.056. Dauchy, X., V. Boiteux, A. Colin, J. Hemard, C. Bach, C. Rosin, and J.F. Munoz. 2019. Deep seepage of per- and polyfluoroalkyl substances through the soil of a firefighter training site and subsequent groundwater contamination. Chemosphere 214:729-737. https://www.sciencedirect.com/science/article/abs/pii/S0045653518318514. D'Hollander, W., D. Herzke, S. Huber, J. Hajslova, J. Pulkrabova, G. Brambilla, S.P. De Filippis, L. Bervoets, and P. de Voogt. 2015. Occurrence of perfluorinated alkylated substances in cereals, salt, sweets and fruit items collected in four European countries. Chemosphere 129:179-185. http://dx.doi.Org/10.1016/i.chemosphere.2014.10.011. Domingo, J.L., I.E. Jogsten, U. Eriksson, I. Martorell, G. Perello, M. Nadal, and B. van Bavel. 2012. Human dietary exposure to perfluoroalkyl substances in Catalonia, Spain. Food Chemistry 135:1575-1582. http://dx.doi.Org/10.1016/i.foodchem.2012.06.054. 39 ------- Dreyer, A., S. Thuens, T. Kirchgeorg, and M. Radk. 2012. Ombrotrophic peat bogs are not suited as natural archives to investigate the historical atmospheric deposition of perfluoroalkyl substances. Environmental Science & Technology 46:7512-7519. http://dx.doi.org/10.1021/es204175y. Eberle, D., R. Ball, and T.B. Boving. 2017. Impact of ISCO treatment on PFAA co-contaminants at a former fire training area. Environmental Science & Technology 51:5127-5136. http://dx.doi.org/10.1021/acs.est.6b06591. ECHA (European Chemicals Agency). 2019. Support Document for Identification of Perfluorobutane Sulfonic Acid and its Salts as Substances of Very High Concern Because of Their Hazardous Properties Which Cause Probable Serious Effects to Human Health and the Environment Which Give Rise to An Equivalent Level of Concern to Those ofCMR and PBT/vPvB Substances (Article 57F). Accessed February 2024. https://echa.europa.eu/documents/10162/891ab33d-d263-cc4b-0f2d-d84cfb7f424a. EFSA (European Food Safety Authority). 2012. Perfluoroalkylated substances in food: Occurrence and dietary exposure. EFSA Journal 10:2743. Accessed February 2024. https://doi.Org/10.2903/i.efsa.2012.2743. EPA (Environmental Protection Agency). 1986. Guidelines for the Health Risk Assessment of Chemical Mixtures. EPA/630/R-98/002. EPA, Risk Assessment Forum, Washington, DC. https://www.epa.gov/risk/guidelines-health-risk-assessment-chemical-mixtures. EPA (Environmental Protection Agency). 2000a. Methodology for Deriving Ambient Water Quality Criteria for the Protection of Human Health (2000). EPA-822-B-00-004. EPA, Office of Water, Office of Science and Technology, Washington, DC. Accessed January 2024. https://www.epa.gov/sites/default/files/2018-10/documents/methodology-wqc- protection-hh-2000.pdf. EPA (Environmental Protection Agency). 2000b. Methodology for Deriving Ambient Water Quality Criteria for the Protection of Human Health (2000), Technical Support Document Vol. 1: Risk Assessment. EPA-822-B-00-005. EPA, Office of Water, Office of Science and Technology, Washington, DC. Accessed January 2024. https://www.epa.gov/sites/default/files/2018-12/documents/methodology-wqc- protection-hh-2000-volumel.pdf. EPA (Environmental Protection Agency). EPA. 2000c. Revisions to the Methodology for Deriving Ambient Water Quality Criteria for the Protection of Human Health (2000); Notice. Federal Register, Nov. 3, 2000, 65:66444. https://www.govinfo.gov/content/pkg/FR-2000-ll-03/pdf/0Q-27924.pdf EPA (Environmental Protection Agency). 2000d. Supplementary Guidance for Conducting Health Risk Assessment of Chemical Mixtures. EPA/630/R-00/002. EPA, Risk Assessment Forum, Washington, DC. https://cfpub.epa.gov/ncea/risk/recordisplay.cfm?deid=20533. 40 ------- EPA (Environmental Protection Agency). 2002. A Review of the Reference Dose and Reference Concentration Processes. EPA/630/P-02/002F. EPA, Risk Assessment Forum, Washington, DC. Accessed March 2024. https://www.epa.gov/sites/clefault/files/2014-12/clocuments/rfcl-final.pclf. EPA (Environmental Protection Agency). 2003. Methodology for Deriving Ambient Water Quality Criteria for the Protection of Human Health (2000), Technical Support Document Volume 2: Development of National Bioaccumulation Factors. EPA-822-R-03-030. EPA, Office of Water, Office of Science and Technology, Washington, DC. Accessed January 2024. https://www.epa.gov/sites/default/files/2018-10/documents/methodology-wqc- protection-hh-2000-volume2.pdf. EPA (Environmental Protection Agency). 2005. Guidelines for Carcinogen Risk Assessment. EPA- 630-P-03-001F. EPA, Risk Assessment Forum, Washington, DC. Accessed January 2024. https://www.epa.gov/sites/default/files/2013- 09/documents/cancer guidelines final 3-25-05.pdf. EPA (Environmental Protection Agency). 2007. Perfluoroalkyl Sulfonates; Significant New Use Rule. EPA, Office of Pollution Prevention and Toxics, Washington, DC. Federal Register, Oct. 9, 2007, 72:57222. Accessed February 2024. https://www.govinfo.gov/content/pkg/FR-2007-10-09/pdf/E7-19828.pdf. EPA (Environmental Protection Agency). 2009. The National Study of Chemical Residues in Lake Fish Tissue. EPA-823-R-09-006. EPA, Office of Water, Office of Science and Technology, Washington, DC. Accessed January 2024. https://www.epa.gov/sites/default/files/2015-07/documents/fish-studv-summarv- 2009.pdf. EPA (Environmental Protection Agency). 2011. Body Weight Studies. Chapter 8 in Exposure Factors Handbook. EPA/600/R-09/052F. EPA, National Center for Environmental Assessment, Office of Research and Development, Washington, DC. Accessed January 2024. https://www.epa.gov/sites/default/files/2015-09/documents/efh-chapter08.pdf. EPA (Environmental Protection Agency). 2014a. Framework for Human Health Risk Assessment to Inform Decision Making. EPA/100/R-14/001. EPA, Office of the Science Advisor, Risk Assessment Forum. Accessed January 2024. https://www.epa.gov/sites/default/files/2014-12/documents/hhra-framework-final- 2014.pdf. EPA (Environmental Protection Agency). 2014b. Estimated Fish Consumption Rates for the U.S. Population and Selected Subpopulations (NHANES 2003-2010). EPA-820-R-14-002. EPA. Accessed January 2024. https://www.epa.gov/sites/default/files/2015- 01/documents/fish-consumption-rates-2014.pdf. 41 ------- EPA (Environmental Protection Agency). 2015. Human Health Ambient Water Quality Criteria: 2015 Update. EPA 820-F-15-001. EPA, Office of Water, Washington, DC. Accessed April 2024. https://www.epa.gov/sites/clefault/files/2015-10/clocuments/human-health- 2015-update-factsheet.pdf. EPA (Environmental Protection Agency). 2016. Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA). EPA-822-R-16-005. EPA, Office of Water, Washington, DC. Accessed January 2024. https://www.epa.gov/sites/default/files/2016- 05/documents/pfoa health advisory final-plain.pdf. EPA (Environmental Protection Agency). 2018. Basic Information on PFAS. https://19ianuary2021snapshot.epa.gov/pfas/basic-information-pfas .html. EPA (Environmental Protection Agency). 2019. Update for Chapter 3 of the Exposure Factors Handbook, Ingestion of Water and Other Select Liquids. EPA/600/R-18/259F. EPA, National Center for Environmental Assessment, Office of Research and Development, Washington, DC. Accessed January 2024. https://www.epa.gov/sites/default/files/2019-02/documents/efh - chapter 3 update.pdf. EPA (Environmental Protection Agency). 2020. National Rivers and Streams Assessment 2013- 2014 Technical Support Document. EPA 843-R-19-001. EPA, Office of Water, Office of Wetlands, Oceans and Watersheds and Office of Research and Development, Washington, DC. Accessed January 2024. https://www.epa.gov/sites/default/files/2020-12/documents/nrsa 2013- 14 final tsd 12-15-2020.pdf. EPA (Environmental Protection Agency). 2021a. Provisional Peer-Reviewed Toxicity Values for Perfluorobutane Sulfonic Acid (PFBS) and Related Compound Potassium Perfluorobutane Sulfonate. EPA/690/R-21/001. EPA, Office of Research and Development, Center for Public Health and Environmental Assessment, Cincinnati, OH. Accessed February 2024. https://cfpub.epa.gov/ncea/pprtv/recordisplay.cfm?deid=350061. EPA (Environmental Protection Agency). 2021b. Human Health Toxicity Values for Perfluorobutane Sulfonic Acid (CASRN 375-73-5) and Related Compound Potassium Perfluorobutane Sulfonate (CASRN 29420-49-3). EPA-600-R-20-345F. EPA, Office of Research and Development, Washington, DC. Accessed February 2024. https://www.epa.gov/pfas/learn-about-human-health-toxicity-assessment-pfbs. 42 ------- EPA (Environmental Protection Agency). 2021c. National Coastal Condition Assessment 2015 Technical Support Document. EPA-841-R-20-002. EPA, Office of Water, Office of Wetlands Oceans and Watersheds and EPA Office of Research and Development, Washington, DC. Accessed February 2024. https://www.epa.gov/svstem/files/documents/2022- 07/NCCA%202015%20TSD%20FI NAL.20210901.pdf. EPA (Environmental Protection Agency). 2022a. National Lakes Assessment: The Third Collaborative Survey of Lakes in the United States. EPA 841-R-22-002. EPA, Office of Water and Office of Research and Development, Washington, DC. Accessed March 2024. https://nationallakesassessment.epa.gov/webreport. EPA (Environmental Protection Agency). 2022b. Transmittal of the Science Advisory Board Report titled, "Review of EPA's Analyses to Support EPA's National Primary Drinking Water Rulemaking for PFAS." EPA-22-008. https://sab.epa.gOv/ords/sab/r/sab apex/sab/advisoryreports. EPA (Environmental Protection Agency). 2023. National Rivers and Streams Assessment 2018- 2019 Technical Support Document. EPA 841-R-22-005. EPA, Office of Water and Office of Research and Development. Washington, DC. Accessed April 2024. https://www.epa.gov/national-aquatic-resource-surveys/nrsa. EPA (Environmental Protection Agency). 2024a. Maximum Contaminant Level Goals (MCLGs) for Three Individual Per- and Polyfluoroalkyl Substances (PFAS) and a Mixture of Four PFAS. EPA-815-R-24-004. EPA, Office of Water, Washington, DC. Accessed April 2024. https://www.epa.gov/system/files/documents/2024-04/pfas-hi-mclg final508.pdf. EPA (Environmental Protection Agency). 2024b. Per- and Polyfluoroalkyl Substances (PFAS) Occurrence and Contaminant Background Support Document for the Fianl PFAS National Primary Drinking Water Regulation. EPA 815-R-24-013. EPA, Office of Water, Washington, DC. Accessed April 2024. https://www.epa.gov/system/files/documents/2024-04/updated-technical-support- document-on-pfas-occurrence final508.pdf. EPA (Environmental Protection Agency). 2024d. Final Freshwater Aquatic Life Ambient Water Quality Criteria and Acute Saltwater Aquatic Life Benchmark for Perfluorooctanoic Acid (PFOA). EPA-842-R-24-002. EPA, Office of Water, Office of Science and Technology, Washington, DC. https://www.epa.gov/system/files/documents/2024-09/pfoa-report-2024.pdf. EPA (Environmental Protection Agency). 2024e. Final Freshwater Aquatic Life Ambient Water Quality Criteria and Acute Saltwater Aquatic Life Benchmark for Perfluorooctane Sulfonate (PFOS). EPA-842-R-24-003. EPA, Office of Water, Office of Science and Technology, Washington, DC. https://www.epa.gov/system/files/documents/2024-09/pfos-report-2024.pdf. 43 ------- EPA (Environmental Protection Agency). 2024f. Integrated Risk Information System. EPA, Office of Research and Development, Washington, DC. Accessed January 2024. https://www.epa.gov/iris. EPA (Environmental Protection Agency). 2024g. Provisional Peer-Reviewed Toxicity Values (PPRTVs). EPA, Office of Research and Development, Center for Public Health and Environmental Assessment, Washington, DC. Accessed January 2024. https://www.epa.gov/pprtv. EPA (Environmental Protection Agency). 2024h. Risk Assessment. EPA, Office of Research and Development, Washington, DC. Accessed January 2024. https://www.epa.gov/risk. EPA (Environmental Protection Agency). 2024i. Pesticide Chemical Search. EPA, Office of Pesticide Programs, Washington, DC. Accessed January 2024. https://ordspub.epa .gov/ords/pesticides/f?p=chemicalsearch:l. EPA (Environmental Protection Agency). 2024j. TSCA Chemical Substance Inventory. EPA, Office of Pollution Prevention and Toxics, Washington, DC. Accessed January 2024. https://www.epa.gov/tsca-inventory. EPA (Environmental Protection Agency). 2024k. EPA Non-Regulatory Health-Based Drinking Water Levels. EPA, Office of Water, Washington, DC. Accessed January 2024. https://www.epa.gov/sdwa/epa-non-regulatory-health-based-drinking-water-levels. EPA (Environmental Protection Agency). 20241. Toxics Release Inventory (TRI) Explorer—Release Reports: Release Chemical Report Page. 2021 Updated dataset (released October 2023). EPA, Washington, DC. Accessed February 2024. https://enviro.epa.gov/triexplorer/tri release.chemical. EPA (Environmental Protection Agency). 2024m. Initial List of Hazardous Air Pollutants with Modifications. EPA, Air Toxics Assessment Group, Research Triangle Park, NC. Accessed February 2024. https://www.epa.goV/haps/initial-list-hazardous-air-pollutants-modifications#mods. EPA (Environmental Protection Agency). 2024n. Method 1633. Analysis of Per-and Polyfluoroalkyl Substances (PFAS) in Aqueous, Solid, Biosolids, and Tissue Samples by LC- MS/MS. EPA 821-R-24-001. EPA, Office of Water, Washington, DC. https://www.epa.gov/cwa-methods/cwa-analvtical-methods-and-polvfluorinated-alkyl- substances-pfas. EPA (Environmental Protection Agency). 2024o. Final Framework for Estimating Noncancer Health Risks Associated with Mixtures of Per-and Polyfluoroalkyl Substances (PFAS). EPA- 815-R-24-003. EPA, Office of Water, Washington, DC. https://www.epa.gov/svstem/files/documents/2024-04/final-pfas-mix-framework- 3.25.24 final-508.pdf. 44 ------- EPA (Environmental Protection Agency). 2024p. Responses to Public Comments on Per- and Polyfluoroalkyl Substances (PFAS) National Primary Drinking Water Regulation Rulemaking. EPA-815-R-24-005. EPA, Office of Water, Washington, DC. https://www.epa.gov/svstem/files/documents/2024-04/pfas-comment-response- document final-508 v2.pdf. Ericson, I., M. Nadal, B. van Bavel, G. Lindstrom, and J.L. Domingo. 2008. Levels of perfluorochemicals in water samples from Catalonia, Spain: Is drinking water a significant contribution to human exposure? Environmental Science and Pollution Research 15:614-619. http://dx.doi.org/10.1007/sll356-008-004Q-l. Eriksson, U., A. Karrman, A. Rotander, B. Mikkelsen, and M. Dam. 2013. Perfluoroalkyl substances (PFASs) in food and water from Faroe Islands. Environmental Science and Pollution Research 20:7940-7948. http://dx.doi.org/10.1007/sll356-013-170Q-3. Eschauzier, C., E. Beerendonk, P. Scholte-Veenendaal, and P. De Voogt. 2012. Impact of treatment processes on the removal of perfluoroalkyl acids from the drinking water production chain. Environmental Science & Technology 46:1708-1715. http://dx.doi.org/10.1021/es2Q1662b. Eschauzier, C., M. Hoppe, M. Schlummer, and P. de Voogt. 2013. Presence and sources of anthropogenic perfluoroalkyl acids in high-consumption tap-water based beverages. Chemosphere 90:36-41. http://dx.doi.Org/10.1016/i.chemosphere.2012.06.070. Favreau, P., C. Poncioni-Rothlisberger, B.J. Place, H. Bouchex-Bellomie, A. Weber, J. Tremp, J.A. Field, and M. Kohler. 2016. Multianalyte profiling of per- and polyfluoroalkyl substances (PFASs) in liquid commercial products. Chemosphere 171:491-501. http://dx.doi.Org/10.1016/i.chemosphere.2016.ll.127. FDA (Food and Drug Administration). 2018. Analytical Results for PFAS in 2018 Produce Sampling (Parts Per Trillion). U.S. Department of Health and Human Services, FDA, Silver Spring, MD. Accessed January 2024. https://www.fda.gov/media/127848/download. FDA (Food and Drug Administration). 2020a. Analytical Results for PFAS in 2019 Total Diet Study Sampling (Parts Per Trillion)—Dataset 1. U.S. Department of Health and Human Services, FDA, Silver Spring, MD. Accessed January 2024. https://www.fda.gov/media/127852/download. FDA (Food and Drug Administration). 2020b. Analytical Results for PFAS in 2019 Total Diet Study Sampling (Parts Per Trillion)—Dataset 2. U.S. Department of Health and Human Services, FDA, Silver Spring, MD. Accessed January 2024. https://www.fda.gov/media/133693/download. 45 ------- FDA (Food and Drug Administration). 2021a. Analytical Results for PFAS in 2021 Total Diet Study Sampling (Parts Per Trillion)—Dataset 3. U.S. Department of Health and Human Services, FDA, Silver Spring, MD. Accessed January 2024. https://www.fda.gov/media/150338/download. FDA (Food and Drug Administration). 2021b. Analytical Results for PFAS in 2021 Total Diet Study Sampling (Parts Per Trillion)—Dataset 4. U.S. Department of Health and Human Services, FDA, Silver Spring, MD. Accessed January 2024. https://www.fda.gov/media/151574/download. FDA (Food and Drug Administration). 2021c. Analytical Results for PFAS in 2018-2021 Dairy Farm Sampling (Parts Per Trillion). U.S. Department of Health and Human Services, FDA, Silver Spring, MD. Accessed January 2024. https://www.fda.gov/media/127850/download. FDA (Food and Drug Administration). 2023. Analytical Results of Testing Food for PFAS from Environmental Contamination. FDA, Silver Spring, MD. Accessed February 2024. https://www.fda.gov/food/chemical-contaminants-food/analytical-results-testing-food- pfas-environmental-contamination. Feng, X., X. Cao, S. Zhao, X. Wang, X. Hua, L. Chen, and L. Chen. 2017. Exposure of pregnant mice to perfluorobutanesulfonate causes hypothyroxinemia and developmental abnormalities in female offspring. ToxicologicalSciences 155(2):409-419. https://doi.org/10.1093/toxsci/kfw219. Fraser, A.J., T.F. Webster, D.J. Watkins, M.J. Strynar, K. Kato, A.M. Calafat, V.M. Vieira, and M.D. McClean. 2013. Polyfluorinated compounds in dust from homes, offices, and vehicles as predictors of concentrations in office workers' serum. Environment International 60:128-136. https://pubmed.ncbi.nlm.nih.gov/24041736/. Galloway, J.E., A.V.P. Moreno, A.B. Lindstrom, M.J. Strynar, S. Newton, A.A. May, and L.K. Weavers. 2020. Evidence of air dispersion: HFPO-DA and PFOA in Ohio and West Virginia surface water near a fluoropolymer production facility. Environmental Science & Technology 54(12):7175-7184. https://doi.org/10.1021/acs.est.9b07384. Gebbink, W.A., L. van Asseldonk, and S.P.J, van Leeuwen. 2017. Presence of emerging per- and polyfluoroalkyl substances (PFASs) in river and drinking water near a fluorochemical production plant in the Netherlands. Environmental Science & Technology 51:11057- 11065. http://dx.doi.org/10.1021/acs.est.7b02488. Gobelius, L., J. Hedlund, W. Diirig, R. Troger, K. Lilja, K. Wiberg, and L. Ahrens. 2018. Per- and polyfluoroalkyl substances in Swedish groundwater and surface water: Implications for environmental quality standards and drinking water guidelines. Environmental Science & Technology 52:4340-4349. http://dx.doi.org/10.1021/acs.est.7b05718. 46 ------- Gremmel, C., T. Fromel, and T.P. Knepper. 2016. Systematic determination of perfluoroalkyl and polyfluoroalkyl substances (PFASs) in outdoor jackets. Chemosphere 160:173-180. http://dx.doi.Org/10.1016/i.chemosphere.2016.06.043. Groffen, T., M. Eens, and L. Bervoets. 2019. Do concentrations of perfluoroalkylated acids (PFAAs) in isopods reflect concentrations in soil and songbirds? A study using a distance gradient from a fluorochemical plant. Science of the Total Environment 657:111-123. https://pubmed.ncbi.nlm.nih.gov/30537574/. Gr0nnestad, R., B.P. Vazquez, A. Arukwe, V.L.B. Jaspers, B.M. Jenssen, M. Karimi, J.L. Lyche, and A. Kr0kje. 2019. Levels, patterns, and biomagnification potential of perfluoroalkyl substances in a terrestrial food chain in a Nordic skiing area. Environmental Science & Technology 53:13390-13397. https://pubmed.ncbi.nlm.nih.gov/31691564/. Hale, S.E., H.P. Arp, G.A. Slinde, E.J. Wade, K. Bj0rseth, G.D. Breedveld, B.F. Straith, K.G. Moe, M. Jartun, and A. H0isaeter. 2017. Sorbent amendment as a remediation strategy to reduce PFAS mobility and leaching in a contaminated sandy soil from a Norwegian firefighting training facility. Chemosphere 171:9-18. https://www.sciencedirect.eom/science/a rticle/abs/pii/S0045653516317775. Harrad, S., N. Wemken, D.S. Drage, M.A.E. Abdallah, and A.M. Coggins. 2019. Perfluoroalkyl substances in drinking water, indoor air and dust from Ireland: Implications for human exposure. Environmental Science & Technology 53:13449-13457. http://dx.doi.org/10.1021/acs.est.9b04604. Harrad, S., D.S. Drage, M. Sharkey, and H. Berresheim. 2020. Perfluoroalkyl substances and brominated flame retardants in landfill-related air, soil, and groundwater from Ireland. Science of the Total Environment 705:135834. http://dx.doi.Org/10.1016/i.scitotenv.2019.135834. Haug, L.S., S. Huber, M. Schlabach, G. Becher, and C. Thomsen. 2011. Investigation on per- and polyfluorinated compounds in paired samples of house dust and indoor air from Norwegian homes. Environmental Science & Technology 45:7991-7998. http://dx.doi.org/10.1021/eslQ3456h. HC (Health Canada). 2023. Health Canada. HC, Ottawa, Ontario, Canada. Accessed January 2024. https://www.canada.ca/en/health-canada.html. Hlouskova, V., P. Hradkova, J. Poustka, G. Brambilla, S.P. De Filipps, W. D'Hollander, L. Bervoets, D. Herzke, S. Huber, P. de Voogt, and J. Pulkrabova. 2013. Occurrence of perfluoroalkyl substances (PFASs) in various food items of animal origin collected in four European countries. Food Additives and Contaminants: Part A 30:1918-1932. http://dx.doi.org/10.1080/19440049.2Q13.837585. 47 ------- Huber, S., L.S. Haug, and M. Schlabach. 2011. Per- and polyfluorinated compounds in house dust and indoor air from northern Norway—a pilot study. Chemosphere 84:1686-1693. http://dx.doi.Org/10.1016/i.chemosphere.2011.04.075. Jogsten, I.E., M. Nadal, B. van Bavel, G. Lindstrom, and J.L. Domingo. 2012. Per- and polyfluorinated compounds (PFCs) in house dust and indoor air in Catalonia, Spain: Implications for human exposure. Environment International 39:172-180. http://dx.doi.Org/10.1016/i.envint.2011.09.004. Karrman, A., I. Ericson, B. van Bavel, P.O. Darnerud, M. Aune, A. Glynn, S. Lignell, and G. Lindstrom. 2007. Exposure of perfluorinated chemicals through lactation: Levels of matched human milk and serum and a temporal trend, 1996-2004, in Sweden. Environmental Health Perspectives 115:226-230. http://dx.doi.org/10.1289/ehp.9491. Karrman, A., J.L. Domingo, X. Llebaria, M. Nadal, E. Bigas, B. van Bavel, and G. Lindstrom. 2010. Biomonitoring perfluorinated compounds in Catalonia, Spain: Concentrations and trends in human liver and milk samples. Environmental Science and Pollution Research 17:750- 758. http://dx.doi.org/10.1007/sll356-009-Q178-5. Kato, K., A.M. Calafat, and L.L. Needham. 2009. Polyfluoroalkyl chemicals in house dust. Environmental Research 109:518-523. http://dx.doi.Org/10.1016/i.envres.2009.01.005. Kotthoff, M., J. Miiller, H. Jurling, M. Schlummer, and D. Fiedler. 2015. Perfluoroalkyl and polyfluoroalkyl substances in consumer products. Environmental Science and Pollution Research 22:14546-14559. http://dx.doi.org/10.1007/sll356-015-42Q2-7. Labadie, P., and M. Chevreuil. 2011. Biogeochemical dynamics of perfluorinated alkyl acids and sulfonates in the River Seine (Paris, France) under contrasting hydrological conditions. Environmental Pollution 159:3634-3639. http://dx.doi.Org/10.1016/i.envpol.2011.07.028. Lankova, D., O. Lacina, J. Pulkrabova, and J. Hajslova. 2013. The determination of perfluoroalkyl substances, brominated flame retardants and their metabolites in human breast milk and infant formula. Talanta 117:318-325. http://dx.doi.Org/10.1016/i.talanta.2013.08.040. Lasier, P.J., J.W. Washington, S.M. Hassan, and T.M. Jenkins. 2011. Perfluorinated chemicals in surface waters and sediments from northwest Georgia, USA, and their bioaccumulation in Lumbriculus variegatus. Environmental Toxicology and Chemistry 30:2194-2201. http://dx.doi.org/10.10Q2/etc.622. Lescord, G.L., K.A. Kidd, A.O. De Silva, M. Williamson, C. Spencer, X. Wang, and D.C. Muir. 2015. Perfluorinated and polyfluorinated compounds in lake food webs from the Canadian high Arctic. Environmental Science and Technology 49:2694-2702. http://dx.doi.org/10.1021/es5Q48649. 48 ------- Loos, R., S. Tavazzi, G. Mariani, G. Suurkuusk, B. Paracchini, and G. Umlauf. 2017. Analysis of emerging organic contaminants in water, fish and suspended particulate matter (SPM) in the Joint Danube Survey using solid-phase extraction followed by UHPLC-MS-MS and GC-MS analysis. Science of the Total Environment 607-608:1201-1212. http://dx.doi.Org/10.1016/i.scitotenv.2017.07.039. Lindstrom, A.B., M.J. Strynar, A.D. Delinsky, S.F. Nakayama, L. McMillan, E.L. Libelo, M. Neill, and L. Thomas. 2011. Application of WWTP biosolids and resulting perfluorinated compound contamination of surface and well water in Decatur, Alabama, USA. Environmental Science & Technology 45:8015-8021. http://dx.doi.org/10.1021/eslQ39425. Liu, X., Z. Guo, K.A. Krebs, R.H. Pope, and N.F. Roache. 2014. Concentrations and trends of perfluorinated chemicals in potential indoor sources from 2007 through 2011 in the US. Chemosphere 98:51-57. http://dx.doi.Org/10.1016/i.chemosphere.2013.10.001. Loi, E.I.H., L.W.Y. Yeung, S. Taniyasu, P.K.S. Lam, K. Kannan, and N. Yamashita. 2011. Trophic magnification of poly- and perfluorinated compounds in a subtropical food web. Environmental Science & Technology 45:5506-5513. https://pubmed.ncbi.nlm.nih.gov/21644538/. Lorenzo, M., J. Campo, M. Farre, F. Perez, Y. Pico, and D. Barcelo. 2015. Perfluoroalkyl substances in the Ebro and Guadalquivir river basins (Spain). Science of the Total Environment 540:191-199. http://dx.doi.Org/10.1016/i.scitotenv.2015.07.045. Mejia-Avendano, S., G. Munoz, S. Vo Duy, M. Desrosiers, P. Benoit, S. Sauve, and J. Liu. 2017. Novel fluoroalkylated surfactants in soils following firefighting foam deployment during the Lac-Megantic railway accident. Environmental Science & Technology 51:8313-8323. http://dx.doi.org/10.1021/acs.est.7b02028. Moller, A., L. Ahrens, R. Surm, J. Westerveld, F. van der Wielen, R. Ebinghaus, and P. de Voogt. 2010. Distribution and sources of polyfluoroalkyl substances (PFAS) in the River Rhine watershed. Environmental Pollution 158:3243-3250. http://dx.doi.Org/10.1016/i.envpol.2010.07.019. Munoz, G., L.C. Fechner, E. Geneste, P. Pardon, H. Budzinski, and P. Labadie. 2016. Spatio- temporal dynamics of per and polyfluoroalkyl substances (PFASs) and transfer to periphytic biofilm in an urban river: Case-study on the River Seine. Environmental Science and Pollution Research 25:23574-23582. http://dx.doi.org/10.1007/sll356-016-8Q51-9. Mussabek, D., L. Ahrens, K.M. Persson, and R. Berndtsson. 2019. Temporal trends and sediment-water partitioning of per- and polyfluoroalkyl substances (PFAS) in lake sediment. Chemosphere 227:624-629. http://dx.doi.Org/10.1016/i.chemosphere.2019.04.074. 49 ------- Nakayama, S., M.J. Strynar, L. Helfant, P. Egeghy, X. Ye, and A.B. Lindstrom. 2007. Perfluorinated compounds in the Cape Fear Drainage Basin in North Carolina. Environmental Science & Technology 41:5271-5276. http://dx.doi.org/10.1021/es070792y. Nakayama, S.F., M.J. Strynar, J.L. Reiner, A.D. Delinsky, and A.B. Lindstrom. 2010. Determination of perfluorinated compounds in the Upper Mississippi River Basin. Environmental Science & Technology 44:4103-4109. http://dx.doi.org/10.1021/esl00382z. Newsted, J.L., R. Holem, G. Hohenstein, C. Lange, M. Ellefson, W. Reagen, and S. Wolf. 2017. Spatial and temporal trends of poly- and perfluoroalkyl substances in fish fillets and water collected from pool 2 of the Upper Mississippi River. Environmental Toxicology and Chemistry 36:3138-3147. http://dx.doi.org/10.10Q2/etc.3891. Newton, S., R. McMahen, J.A. Stoeckel, M. Chislock, A. Lindstrom, and M. Strynar. 2017. Novel polyfluorinated compounds identified using high resolution mass spectrometry downstream of manufacturing facilities near Decatur, Alabama. Environmental Science & Technology 51:1544-1552. http://dx.doi.org/10.1021/acs.est.6b05330. Nickerson, A., A.C. Maizel, P.R. Kulkarni, D.T. Adamson, J.J. Kornuc, and C.P. Higgins. 2020. Enhanced extraction of AFFF-associated PFASs from source zone soils. Environmental Science & Technology 54:4952-4962. http://dx.doi.org/10.1021/acs.est.0cQ0792. NOAA (National Oceanic and Atmospheric Administration). 2024. NOAA's National Status and Trends Data Page. NCCOS Data Collections. U.S. Department of Commerce, NOAA, National Centers for Coastal Ocean Science, Silver Spring, MD. Accessed February 2024. https://products.coastalscience.noaa.gov/nsandt data/data.aspx. Nyberg, E., R. Awad, A. Bignert, C. Ek, G. Sallsten, and J.P. Benskin. 2018. Inter-individual, inter- city, and temporal trends of per- and polyfluoroalkyl substances in human milk from Swedish mothers between 1972 and 2016. Environmental Science: Process Impacts 20:1136-1147. http://dx.doi.org/10.1039/c8em00174i. Pan, Y., H. Zhang, Q. Cui, N. Sheng, L.W.Y. Yeung, Y. Sun, Y. Guo, and J. Dai. 2018. Worldwide distribution of novel perfluoroether carboxylic and sulfonic acids in surface water. Environmental Science and Technology 52:7621-7629. https://pubs.acs.org/doi/10.1021/acs.est.8b00829. Papadopoulou, E., S. Poothong, J. Koekkoek, L. Lucattini, J.A. Padilla-Sanchez, M. Haugen, D. Herzke, S. Valdersnes, A. Maage, I.T. Cousins, P.E.G. Leonards, and L.S. Haug. 2017. Estimating human exposure to perfluoroalkyl acids via solid food and drinks: Implementation and comparison of different dietary assessment methods. Environmental Research 158:269-276. http://dx.doi.Org/10.1016/i.envres.2017.06.011. 50 ------- Perez, F., M. Llorca, M. Kock-Schulmeyer, B. Skrbic, L.S. Oliveira, K. da Boit Martinello, N.A. Al- Dhabi, I. Antic, M. Farre, and D. Barcelo. 2014. Assessment of perfluoroalkyl substances in food items at global scale. Environmental Research 135:181-189. http://dx.doi.Org/10.1016/i.envres.2014.08.004. Post, G.B., J.B. Louis, R.L. Lippincott, and N.A. Procopio. 2013. Occurrence of perfluorinated compounds in raw water from New Jersey public drinking water systems. Environmental Science & Technology 47:13266-13275. http://dx.doi.org/10.1021/es4Q2884x. Procopio, N.A., R. Karl, S.M. Goodrow, J. Maggio, J.B. Louis, and T.B. Atherholt. 2017. Occurrence and source identification of perfluoroalkyl acids (PFAAs) in the Metedeconk River Watershed, New Jersey. Environmental Science and Pollution Research 24:27125- 27135. http://dx.doi.org/10.1007/sll356-017-03Q9-3. Rostkowski, P., S. Taniyasu, N. Yamashita, J.J. Falandysz, t. Zegarowski, A. Chojnacka, K. Pazdro, and J. Falandysz. 2009. Survey of perfluorinated compounds (PFCs) in surface waters of Poland. Journal of Environmental Science and Health, Part A, Toxic/Hazardous Substances and Environmental Engineering 44:1518-1527. http://dx.doi.org/10.1080/1093452090326333Q. Schaider, L.A., S.A. Balan, A. Blum, D.Q. Andrews, M.J. Strynar, M.E. Dickinson, D.M. Lunderberg, J.R. Lang, and G.F. Peaslee. 2017. Fluorinated compounds in US fast food packaging. Environmental Science & Technology Letters 4:105-111. https://pubs.acs.org/doi/10.1021/acs.estlett.6b00435. Schecter, A., J. Colacino, D. Haffner, K. Patel, M. Opel, O. Papke, and L. Birnbaum. 2010. Perfluorinated compounds, polychlorinated biphenyls, and organochlorine pesticide contamination in composite food samples from Dallas, Texas, USA. Environmental Health Perspectives 118:796-802. http://dx.doi.org/10.1289/ehp.09Q1347. Scher, D.P., J.E. Kelly, C.A. Huset, K.M. Barry, R.W. Hoffbeck, V.L. Yingling, and R.B. Messing. 2018. Occurrence of perfluoroalkyl substances (PFAS) in garden produce at homes with a history of PFAS-contaminated drinking water. Chemosphere 196:548-555. http://dx.doi.Org/10.1016/i.chemosphere.2017.12.179. Scher, D.P., J.E. Kelly, C.A. Huset, K.M. Barry, and V.L. Yingling. 2019. Does soil track-in contribute to house dust concentrations of perfluoroalkyl acids (PFAAs) in areas affected by soil or water contamination? Journal of Exposure Science and Environmental Epidemiology 29:218-226. http://dx.doi.org/10.1038/s41370-018-Q101-6. Schultes, L., R. Vestergren, K. Volkova, E. Westberg, T. Jacobson, and J.P. Benskin. 2018. Per- and polyfluoroalkyl substances and fluorine mass balance in cosmetic products from the Swedish market: Implications for environmental emissions and human exposure. Environmental Science: Process & Impacts 20:1680-1690. http://dx.doi.org/10.1039/c8em00368h. 51 ------- Scordo, C.V.A., L. Checchini, L. Renai, S. Orlandini, M.C. Bruzzoniti, D. Fibbi, L. Mandi, N. Ouazzani, and M. Del Bubba. 2020. Optimization and validation of a method based on QuEChERS extraction and liquid chromatographic-tandem mass spectrometric analysis for the determination of perfluoroalkyl acids in strawberry and olive fruits, as model crops with different matrix characteristics. Journal of Chromatography A 1621:461038. http://dx.doi.Org/10.1016/i.chroma.2020.461038. Shafique, U., S. Schulze, C. Slawik, A. Bohme, A. Paschke, and G. Schuurmann. 2017. Perfluoroalkyl acids in aqueous samples from Germany and Kenya. Environmental Science and Pollution Research 24:11031-11043. http://dx.doi.org/10.1007/sll356-016-7Q76-4. Skaar, J.S., E.M. Raeder, J.L. Lyche, L. Ahrens, and R. Kallenborn. 2019. Elucidation of contamination sources for poly- and perfluoroalkyl substances (PFASs) on Svalbard (Norwegian Arctic). Environmental Science and Pollution Research International 26:7356-7363. https://pubmed.ncbi.nlm.nih.gov/29754295/. Stahl, L.L., B.D. Snyder, A.R. Olsen, T.M. Kincaid, J.B. Wathen, and H.B. McCarty. 2014. Perfluorinated compounds in fish from U.S. urban rivers and the Great Lakes. Science of the Total Environment 499:185-195. https://doi.Org/10.1016/i.scitotenv.2014.07.126. Strynar, M.J., and A.B. Lindstrom. 2008. Perfluorinated compounds in house dust from Ohio and North Carolina, USA. Environmental Science & Technology 42:3751-3756. http://dx.doi.org/10.1021/es7032058. Subedi, B., N. Codru, D.M. Dziewulski, L.R. Wilson, J. Xue, S. Yun, E. Braun-Howland, C. Minihane, and K. Kannan. 2015. A pilot study on the assessment of trace organic contaminants including pharmaceuticals and personal care products from on-site wastewater treatment systems along Skaneateles Lake in New York State, USA. Water Research 72:28-39. https://doi.Org/10.1016/i.watres.2014.10.049. Surma, M., W. Wiczkowski, H. Zielinski, and E. Cieslik. 2015. Determination of selected perfluorinated acids (PFCAs) and perfluorinated sulfonates (PFASs) in food contact materials using LC-MS/MS. Packaging Technology and Science 28:789-799. https://doi.org/10.1002/pts.2140. Surma, M., M. Piskula, W. Wiczkowski, and H. Zielinski. 2017. The perfluoroalkyl carboxylic acids (PFCAs) and perfluoroalkane sulfonates (PFSAs) contamination level in spices. European Food Research and Technology 243:297-307. http://dx.doi.org/10.1007/sQ0217-016-2744-7. Sznajder-Katarzyriska, K., M. Surma, W. Wiczkowski, and E. Cieslik. 2019. The perfluoroalkyl substance (PFAS) contamination level in milk and milk products in Poland. International Dairy Journal 96:73-84. http://dx.doi.Org/10.1016/i.idairvi.2019.04.008. 52 ------- University of Maryland. 2024. What We Eat in America—Food Commodity Intake Database 2005-10. University of Maryland, College Park, MD, and EPA Office of Pesticide Programs, Washington, DC. Accessed February 2024. https://fcid.foodrisk.org/. Valsecchi, S., M. Rusconi, M. Mazzoni, G. Viviano, R. Pagnotta, C. Zaghi, G. Serrini, and S. Polesello. 2015. Occurrence and sources of perfluoroalkyl acids in Italian river basins. Chemosphere 129:126-134. http://dx.doi.Org/10.1016/i.chemosphere.2014.07.044. van der Veen, I., A.C. Hanning, A. Stare, P.E.G. Leonards, J. de Boer, and J.M. Weiss. 2020. The effect of weathering on per- and polyfluoroalkyl substances (PFASs) from durable water repellent (DWR) clothing. Chemosphere 249:126100. http://dx.doi.Org/10.1016/i.chemosphere.2020.126100. Vassiliadou, I., D. Costopoulou, N. Kalogeropoulos, S. Karavoltsos, A. Sakellari, E. Zafeiraki, M. Dassenakis, and L. Leondiadis. 2015. Levels of perfluorinated compounds in raw and cooked Mediterranean finfish and shellfish. Chemosphere 127:117-126. http://dx.doi.Org/10.1016/i.chemosphere.2014.12.081. Vavrous, A., L. Vapenka, J. Sosnovcova, K. Kejlova, K. Vrbik, and D. Jirova. 2016. Method for analysis of 68 organic contaminants in food contact paper using gas and liquid chromatography coupled with tandem mass spectrometry. Food Control 60:221-229. http://dx.doi.Org/10.1016/i.foodcont.2015.07.043. Veillette, J., D.C.G. Muir, D. Antoniades, J.M. Small, C. Spencer, T.N. Loewen, J.A. Babaluk, J.D. Reist, and W.F. Vincent. 2012. Perfluorinated chemicals in meromictic lakes on the northern coast of Ellesmere Island, High Arctic Canada. Arctic 65:245-256. https://www.researchgate.net/publication/263765598 Perfluorinated Chemicals in M eromictic Lakes on the Northern Coast of Ellesmere Island High Arctic Canada. Venkatesan, A.K., and R.U. Halden. 2013. National inventory of perfluoroalkyl substances in archived U.S. biosolids from the 2001 EPA National Sewage Sludge Survey. Journal of Hazardous Materials 252-253:413-418. https://www.sciencedirect.com/science/article/abs/pii/S03Q43894130Q1921. Venkatesan, A.K., and R.U. Halden. 2014. Loss and in situ production of perfluoroalkyl chemicals in outdoor biosolids-soil mesocosms. Environmental Research 132:321-327. http://dx.doi.Org/10.1016/i.envres.2014.04.024. Vestergren, R., D. Herzke T. Wang, and I.T. Cousins. 2015. Are imported consumer products an important diffuse source of PFASs to the Norwegian environment? Environmental Pollution 198:223-230. http://dx.doi.Org/10.1016/i.envpol.2014.12.034. von Ehrenstein, O.S., S.E. Fenton, K. Kato, Z. Kuklenyik, A.M. Calafat, and E.P. Hines. 2009. Polyfluoroalkyl chemicals in the serum and milk of breastfeeding women. Reproductive Toxicology 27:239-245. http://dx.doi.Org/10.1016/i.reprotox.2009.03.001. 53 ------- Wagner, A., B. Raue, H.J. Brauch, E. Worch, and F.T. Lange. 2013. Determination of adsorbable organic fluorine from aqueous environmental samples by adsorption to polystyrene- divinylbenzene based activated carbon and combustion ion chromatography. Journal of Chromatography A 1295:82-89. http://dx.doi.Org/10.1016/i.chroma.2013.04.051. Wilkinson, J.L., P.S. Hooda, J. Swinden, J. Barker, and S. Barton. 2017. Spatial distribution of organic contaminants in three rivers of Southern England bound to suspended particulate material and dissolved in water. Science of the Total Environment 593- 594:487-497. http://dx.doi.Org/10.1016/i.scitotenv.2017.03.167. Wong, F., M. Shoeib, A. Katsoyiannis, S. Eckhardt, A. Stohl, P. Bohlin-Nizzetto, H. Li, P. Fellin, Y. Su, and H. Huang. 2018. Assessing temporal trends and source regions of per-and polyfluoroalkyl substances (PFASs) in air under the Arctic Monitoring and Assessment Programme (AMAP). Atmospheric Environment 172:65-73. https://doi.Org/10.1016/i.atmosenv.2017.10.028. Yeung, L.W.Y., C. Stadey, and S.A. Mabury. 2017. Simultaneous analysis of perfluoroalkyl and polyfluoroalkyl substances including ultrashort-chain C2 and C3 compounds in rain and river water samples by ultra performance convergence chromatography. Journal of Chromatography A 1522:78-85. http://dx.doi.Org/10.1016/i.chroma.2017.09.049. Young, W.M., P. South, T.H. Begley, G.W. Diachenko, and G.O. Noonan. 2012. Determination of perfluorochemicals in cow's milk using liquid chromatography-tandem mass spectrometry. Journal of Agricultural and Food Chemistry 60:1652-1658. http://dx.doi.org/10.1021/if2Q4565x. Young, W.M., P. South, T.H. Begley, and G.O. Noonan. 2013. Determination of perfluorochemicals in fish and shellfish using liquid chromatography-tandem mass spectrometry. Journal of Agricultural and Food Chemistry 61:11166-11172. http://dx.doi.org/10.1021/if4Q3935g. Zafeiraki, E., D. Costopoulou, I. Vassiliadou, E. Bakeas, and L. Leondiadis. 2014. Determination of perfluorinated compounds (PFCs) in various foodstuff packaging materials used in the Greek market. Chemosphere 94:169-176. http://dx.doi.Org/10.1016/i.chemosphere.2013.09.092. Zhang, X., R. Lohmann, C. Dassuncao, X.C. Hu, A.K. Weber, C.D. Vecitis, and E.M. Sunderland. 2016. Source attribution of poly- and perfluoroalkyl substances (PFASs) in surface waters from Rhode Island and the New York metropolitan area. Environmental Science & Technology Letters 3:316-321. http://dx.doi.org/10.1021/acs.estlett.6b00255. Zhao, Z., Z. Xie, J. Tang, R. Sturm, Y. Chen, G. Zhang, and R. Ebinghaus. 2015. Seasonal variations and spatial distributions of perfluoroalkyl substances in the rivers Elbe and lower Weser and the North Sea. Chemosphere 129:118-125. http://dx.doi.Org/10.1016/i.chemosphere.2014.03.050. 54 ------- Zheng, G., B.E. Boor, E. Schreder, and A. Salamova. 2020. Indoor exposure to per- and polyfluoroalkyl substances (PFAS) in the childcare environment. Environmental Pollution 258:113714. http://dx.doi.Org/10.1016/i.envpol.2019.113714. 55 ------- Appendix A: Summary of Supporting Literature for Surface Water Occurrence Table A-l. Compilation of studies describing PFBS occurrence in surface water. Study Location Site Details PFBS Results North America Anderson et al. United States Ten U.S. Air Force DF 80.00%, median (2016) (national) installations with historic AFFF release (range) = 106 (ND- 317,000) ng/L Appleman et al. United States Raw surface waters from DFa 64% (n = 25); range = ND- (2014) (Wisconsin, Oklahoma, 11 sites, some impacted by 47 ng/L Alaska, California, upstream wastewater (MRL = 0.3) Alabama, Colorado, effluent discharge Ohio, Nevada, Minnesota, New Jersey) Bradley et al. United States (Lake Untreated Lake Michigan DF 29%, range = ND-0.5 ng/L (2020) Michigan) water from treatment plant intake (4 sites) Galloway et al. United States (Ohio Rivers and tributaries 58 km DF NR, range3 = ND-28.0 ng/L (2020) and West Virginia; Ohio River Basin) upstream to 130 km downwind of a fluoropolymer production facility, some sample locations potentially impacted by local landfills Lasier et al. United States (Georgia; Upstream (sites 1 and 2) and Upstream (2011) Coosa River downstream (sites 3-8) of a Sites 1 and 2: DF 0% watershed) land-application site where effluents from carpet manufacturers (suspected of producing wastewaters containing perfluorinated chemicals) are processed at a WWTP and the treated WWTP effluent is sprayed onto the site. Site 4 was downstream of a manufacturing facility for latex and polyurethane backing material. Downstream Site 3: DF NR, mean = 205 ng/L Site 4: DF NR, mean = 260 ng/L Site 5: DF NR, mean = 125 ng/L Site 6: DF NR, mean = 134 ng/L Site 7: DF NR, mean = 122 ng/L Site 8: DF NR, mean = 105 ng/L Lescord et al. Canada (Resolute Bay, One lake (Meretta) Meretta: DF NR, (2015) Nunavut) contaminated with runoff from an airport, which is a known source of PFAS; one control lake (9 Mile) mean = 4.9 ng/L 9 Mile: DF NR, mean = 0.07 ng/L 56 ------- Study Location Site Details PFBS Results Lindstrom et al. United States 32 surface water samples DFa 63%, range = ND- (2011) (Alabama) (ponds and streams) from areas with historical land application of fluorochemical industry-impacted biosolids 208 ng/L Nakayama et al. United States (North 80 sampling sites in river DF 62%, mean (range) = 2.58 (2007) Carolina; Cape Fear River Basin) basin; some sites near industrial areas and Fort Bragg and Pope Air Force Base with suspected use of AFFF at the Air Force Base (ND-9.41) ng/L Nakayama et al. United States (Illinois, 88 sampling sites from DF 43%, median (2010) Iowa, Minnesota, Missouri, Wisconsin; Upper Mississippi River Basin and Missouri River Basin) tributaries and streams (range) = 0.71 (ND-84.1) ng/L Newsted et al. United States Upstream and downstream of Upstream: DFa 3%, (2017) (Minnesota; Upper 3M Cottage Grove facility point = 4.2 ng/L Mississippi River Pool outfall, which is a source of Downstream: DFa67%, 2) PFAS range = ND-336.0 ng/L Newton et al. United States (Decatur, 6 sites upstream and 3 sites Upstream: DF 0% (2017) Alabama; Tennessee downstream of Downstream: DFa 100%, River) fluorochemical manufacturing facilities mean3 (range) = 69 (10- 160)ng/L Post et al. United States (New 6 rivers and 6 reservoirs from DF 17%, range = ND-6 ng/L (2013) Jersey) public drinking water system intakes, some sites may include nearby small industrial park and civil- military airport Procopio et al. United States (New Downstream of suspected DFa 5%, range = ND-100 ng/L (2017) Jersey; Metedeconk River Watershed) illicit discharge to soil and groundwater from a manufacturer of industrial fabrics, composites, and elastomers that use or produce products containing PFAAs Subedi et al. United States (New Lake water along the DFa4%(n = 28); single (2015) York; Skaneateles Lake) shoreline of residences that use an enhanced treatment unit for onsite wastewater treatment detection value = 0.26 ng/L 57 ------- Study Location Site Details PFBS Results Veillette et al. Canada (Ellesmere A lake near the northwest DFa 100%, mean (2012) Island, Nunavut) coast with no known sources of PFAS (range) = 0.016 (0.011- 0.024) ng/L Yeung et al. Canada (Ontario; Two water samples at each of Mimico Creek: (2017) Mimico Creek, Rouge River) the sites point = 0.020 ng/L Rouge River: DF 0% Zhang et al. United States (Rhode Rivers and creeks, some DFa 85%, range = ND- (2016) Island, New York Metropolitan Region) sampling locations downstream from industrial activities, airport, textile mills, and WWTP. PFAS are used for water resistant coating in textiles. 6.181 ng/L Europe Ahrens et al. Germany (Elbe River) Sampling sites in Hamburg Hamburg: (2009a) city (sites 16-18) and from Laurenburg to Hamburg (sites 19-24) Dissolved: DFa 100%, mean (range) = 1.6 (1.1-2.5) ng/L Laurenburg to Hamburg: Dissolved: DFa 100%, mean (range) = 1.1 (0.53-1.5) ng/L Ahrens et al. Germany (Elbe River) Sampling locations 53 to DF NR; range of mean (for (2009b) 122 km (sites 1 to 9)c upstream of estuary mouth of Elbe River different locations) = 1.8- 3.4 ng/L Bach et al. France (southern) Upstream and downstream Upstream: DF 0% (2017) from discharge point that receives wastewater from an industrial site with two fluoropolymer manufacturing facilities Downstream: DF 0% Barreca et al. Italy (Lombardia Rivers and streams with no DFa 39%, range = ND- (2020) Region) known fluorochemical sources 16,000 ng/L Boiteux et al. France (national) Rivers; some locations may DF 1%, range = ND-5 ng/L (2012) have upstream industrial sources Boiteux et al. France (northern) River samples from upstream Upstream: DF 0% (2017) and downstream of an industrial WWTP that processes raw sewage from fluorochemical manufacturing facility Downstream: DF 0% Dauchy etal. France (unspecified) Samples collected near 3 sites Site B: DF 0% (2017) (B, C, D) impacted by the use of firefighting foams Site C: DF 0% Site D: DFa 30%, range = ND- 138 ng/L 58 ------- Study Location Site Details PFBS Results Ericson et al. Spain (Tarragona Sampling sites were not Ebro site 1: DF 0% (2008) Province; Ebro River, proximate to known point Ebro site 2: DF 0% Francolf River, Cortiella sources of any fluorochemical Francolf: DF 0% River) facilities Cortiella: DF 0% Eriksson et al. Denmark (Faroe Lakes Leitisvatn, Havnardal, Leitisvatn: DF 0% (2013) Islands) Kornvatn, and A Myranar with no known point sources of any fluorochemical facilities Havnardal Lake: DF 0% Kornvatn Lake: DF 0% A Myranar: DF 0% Eschauzier et al. The Netherlands Downstream of an industrial DFa 100%, mean (range) = 35 (2012) (Amsterdam; Lek Canal, tributary of Rhine River) point source in the German part of the Lower Rhine (31-42)ng/L Gebbink et al. The Netherlands Upstream and downstream of Control sites: DFa 100%, (2017) (Dordrecht) Dordrecht fluorochemical production plant; two control sites mean3 (range) = 17 (12- 22)ng/L Upstream: DFa 100%, mean3 (range) = 19.7 (18-21) ng/L Downstream: DFa 100%, mean3 (range) = 21 (16- 27) ng/L Gobelius et al. Sweden (national) Sampling locations selected DF3 29%, range = ND- (2018) based on potential vicinity of PFAS hot spots and importance as a drinking water source area, some sites include firefighting training sites at airfields and military areas 299 ng/L Labadie and France (Paris; River Urban stretch of the River DF 100%, mean (range) = 1.3 Chevreuil (2011) Seine) Seine during a flood cycle, sampling location under the influence of two urban WWTPs and two major combined sewer overflow outfalls (0.6-2.6) ng/L Loos et al. Austria, Bulgaria, Some sampling locations DF 94%, mean (2017) Croatia, Moldova, Romania, Serbia, Slovakia (Danube River and tributaries) downstream of major cities (range) = 1.6 (ND-3.7) ng/L 59 ------- Study Location Site Details PFBS Results Lorenzo et al. Spain (Guadalquivir Guadalquivir sampling Guadalquivir: DF 8%, mean (2015) River Basin, Ebro River locations included (range) = 10.1 (ND- Basin) downstream of WWTPs, near industrial areas, near a military camp, or through major cities; Ebro sampling locations included nearby ski resorts and downstream of WWTP and industrial areas 228.3) ng/L Ebro: DF 0% Moller et al. Germany (Rhine River Upstream and downstream of Rhine upstream Leverkusen: (2010) watershed) Leverkusen, where effluent of a WWTP treating industrial wastewater was discharged; other major rivers and tributaries DF 100%, mean (range) = 3.19 (0.59-6.58) ng/L Rhine downstream Leverkusen: DF 100%, mean (range) = 45.4 (15.0-118) ng/L River Ruhr: DF 100%, mean (range) = 7.08 (2.87-11.4) ng/L River Moehne: point = 31.1 ng/L Other tributaries: DF 100%, mean (range) = 2.84 (0.22- 6.82) ng/L Munoz et al. France (Seine River) Two sites downstream of DF 70%, range = ND-3.1 ng/L (2016) Greater Paris and one site unaffected by the Greater Paris region Mussabek et al. Sweden (Lulea) Samples from lake and pond Lake: DF NR, mean = 200 ng/L (2019) near a firefighting training facility at the Norrbotten Air Force Wing known to use PFAS-containing AFFF Pond: DF NR, mean = 150 ng/L Rostkowski et Poland (national) Rivers, lakes, and streams in North: DFa 60%, range = ND- al. (2009) northern and southern Poland, some southern locations near chemical industrial activities 10 ng/L South: DFa 73%, range = ND- 16.0 ng/L Shafique et al. Germany (Leipzig, Sampling sites were not Pleil3e-Elster: DF NR, (2017) Pleil3e-Elster River, proximate to known point mean = 1.2 ng/L Saale River, and Elbe sources of any fluorochemical Saale: DF NR, mean = 7.5 ng/L River) facilities Elbe: DF NR, mean = 4.3 ng/L 60 ------- Study Location Site Details PFBS Results Valsecchi et al. Italy (Po River Basin, Two river basins (Po and Po: DFa 56%, range = ND- (2015) Brenta River Basin, Brenta) which receive 30.4 ng/L Adige River Basin, discharges from two chemical Brenta: DFa 100%, mean3 Tevere River Basin, and plants that produce (range) = 707 (23.1- Arno River Basin) fluorinated polymers and intermediates; three river basins (Adige, Tevere, Arno) with no known point sources of any fluorochemical facilities 1,666) ng/L Adige: DFa 20%, range = ND- 4.3 ng/L Tevere: DF 0% Arno: DFa 58%, range = ND- 31.4 ng/L Wagner et al. Germany (Rhine River) Sampling sites were not DFa 100%, meanb (2013) proximate to known point sources of any fluorochemical facilities (rangeb) = 18 (9-26) ng/L Wilkinson et al. England (Greater 50 m upstream and 250 m Upstream: DF NR, (2017) London and southern and 1,000 m downstream mean = 20.4 ng/L England; Hogsmill from WWTP effluent outfalls Downstream 250 m: DF NR, River, Chertsey Bourne mean = 40.3 ng/L River, Blackwater Downstream 1,000 m: DF NR, River) mean = 41.1 ng/L Zhao et al. Germany (Elbe River Some sampling sites near Elbe: DF 100%, mean (2015) and lower Weser River) Hamburg city and industrial plants (range) = 7.4 (0.24-238) ng/L Weser: DF 100%, mean (range) = 1.41 (0.75- 1.85) ng/L Multiple Continents Pan et al. (2018) United States Sampling sites were not DFa 100%, mean (Delaware River) proximate to known point sources of any fluorochemical facilities (range) = 2.19 (0.52- 4.20) ng/L United Kingdom Sampling sites were not DFa 100%, mean (Thames River) proximate to known point sources of any fluorochemical facilities (range) = 5.06 (3.26- 6.75) ng/L Germany and the Sampling sites were not DFa 100%, mean Netherlands (Rhine proximate to known point (range) = 21.9 (0.46-146) ng/L River) sources of any fluorochemical facilities Sweden (Malaren Lake) Sampling sites were not proximate to known point sources of any fluorochemical facilities DFa 100%, mean (range) = 1.43 (0.75- 1.92) ng/L Notes: AFFF = aqueous film-forming foam; DF = detection frequency; km = kilometer; m = meter; ND = not detected; ng/L = nanogram per liter; NR = not reported; PFAA = perfluoroalkyl acid; PFAS = per- and polyfluoroalkyl substances; WWTP = wastewater treatment plant; ng/L = microgram per liter. aThe DF and/or mean was not reported in the study and was calculated in this synthesis. Means were calculated only when DF = 100%. 61 ------- b For Wagner et al. (2013), PFBS concentrations were calculated using the fluorine concentrations reported in Table 4 from the study. c Freshwater locations determined as sites with conductivity < 1.5 milliSiemens/cm. References Ahrens, L., M. Plassmann, Z. Xie, and R. Ebinghaus. 2009a. Determination of polyfluoroalkyl compounds in water and suspended particulate matter in the river Elbe and North Sea, Germany. Frontiers of Environmental Science & Engineering in China 3:152-170. http://dx.doi.org/10.1007/sll783-009-0Q21-8. Ahrens, L., S. Felizeter, R. Sturm, Z. Xie, and R. Ebinghaus. 2009b. Polyfluorinated compounds in waste water treatment plant effluents and surface waters along the River Elbe, Germany. Marine Pollution Bulletin 58:1326-1333. http://dx.doi.Org/10.1016/i.marpolbul.2009.04.028. Anderson, R.H., G.C. Long, R.C. Porter, and J.K. Anderson. 2016. Occurrence of select perfluoroalkyl substances at U.S. Air Force aqueous film-forming foam release sites other than fire-training areas: Field-validation of critical fate and transport properties. Chemosphere 150:678-685. http://dx.doi.Org/10.1016/i.chemosphere.2016.01.014. Appleman, T.D., C.P. Higgins, O. Quinones, B.J. Vanderford, C. Kolstad, J.C. Zeigler-Holady, and E.R. Dickenson. 2014. Treatment of poly- and perfluoroalkyl substances in U.S. full-scale water treatment systems. Water Research 51:246-255. http://dx.doi.Org/10.1016/i.watres.2013.10.067. Bach, C., X. Dauchy, V. Boiteux, A. Colin, J. Hemard, V. Sagres, C. Rosin, and J.F. Munoz. 2017. The impact of two fluoropolymer manufacturing facilities on downstream contamination of a river and drinking water resources with per- and polyfluoroalkyl substances. Environmental Science and Pollution Research 24:4916-4925. http://dx.doi.org/10.1007/sll356-016-8243-3. Barreca, S., M. Busetto, L. Colzani, L. Clerici, V. Marchesi, L. Tremolada, D. Daverio, and P. Dellavedova. 2020. Hyphenated high performance liquid chromatography-tandem mass spectrometry techniques for the determination of perfluorinated alkylated substances in Lombardia region in Italy, profile levels and assessment: One year of monitoring activities during 2018. Separations 7(1):17. http://dx.doi.org/10.3390/separations7010Q17. Boiteux, V., X. Dauchy, C. Rosin, and J.F. Munoz. 2012. National screening study on 10 perfluorinated compounds in raw and treated tap water in France. Archives of Environmental Contamination and Toxicology 63:1-12. http://dx.doi.org/10.1007/sQ0244-012-9754-7. 62 ------- Boiteux, Vv X. Dauchy, C. Bach, A. Colin, J. Hemard, V. Sagres, C. Rosin, and J.F. Munoz. 2017. Concentrations and patterns of perfluoroalkyl and polyfluoroalkyl substances in a river and three drinking water treatment plants near and far from a major production source. Science of the Total Environment 583:393-400. http://dx.doi.Org/10.1016/i.scitotenv.2017.01.079. Bradley, P.M., M. Argos, D.W. Kolpin, S.M. Meppelink, K.M. Romanok, K.L. Smalling, M.J. Focazio, J.M. Allen, J.E. Dietze, M.J. Devito, A.R. Donovan, N. Evans, C.E. Givens, J.L. Gray, C.P. Higgins, M.L. Hladik, L.R. Iwanowicz, C.A. Journey, R.F. Lane, Z.R. Laughrey, K.A. Loftin, R.B. McCleskey, C.A. McDonough, E. Medlock-Kakaley, M.T. Meyer, A.R. Putz, S.D. Richardson, A.E. Stark, C.P. Weis, V.S. Wilson, and A. Zehraoui. 2020. Mixed organic and inorganic tapwater exposures and potential effects in greater Chicago area, USA. Science of the Total Environment 719:137236. http://dx.doi.Org/10.1016/i.scitotenv.2020.137236. Dauchy, X., V. Boiteux, C. Bach, C. Rosin, and J.F. Munoz. 2017. Per- and polyfluoroalkyl substances in firefighting foam concentrates and water samples collected near sites impacted by the use of these foams. Chemosphere 183:53-61. http://dx.doi.Org/10.1016/i.chemosphere.2017.05.056. Ericson, I., M. Nadal, B. van Bavel, G. Lindstrom, and J.L. Domingo. 2008. Levels of perfluorochemicals in water samples from Catalonia, Spain: Is drinking water a significant contribution to human exposure? Environmental Science and Pollution Research 15:614-619. http://dx.doi.org/10.1007/sll356-008-004Q-l. Eriksson, U., A. Karrman, A. Rotander, B. Mikkelsen, and M. Dam. 2013. Perfluoroalkyl substances (PFASs) in food and water from Faroe Islands. Environmental Science and Pollution Research 20:7940-7948. http://dx.doi.org/10.1007/sll356-013-170Q-3. Eschauzier, C., E. Beerendonk, P. Scholte-Veenendaal, and P. De Voogt. 2012. Impact of treatment processes on the removal of perfluoroalkyl acids from the drinking water production chain. Environmental Science & Technology 46:1708-1715. http://dx.doi.org/10.1021/es2Q1662b. Galloway, J.E., A.V.P. Moreno, A.B. Lindstrom, M.J. Strynar, S. Newton, A.A. May, and L.K. Weavers. 2020. Evidence of air dispersion: HFPO-DA and PFOA in Ohio and West Virginia surface water near a fluoropolymer production facility. Environmental Science & Technology 54(12):7175-7184. https://doi.org/10.1021/acs.est.9b07384. Gebbink, W.A., L. van Asseldonk, and S.P.J, van Leeuwen. 2017. Presence of emerging per- and polyfluoroalkyl substances (PFASs) in river and drinking water near a fluorochemical production plant in the Netherlands. Environmental Science & Technology 51:11057- 11065. http://dx.doi.org/10.1021/acs.est.7b02488. 63 ------- Gobelius, L., J. Hedlund, W. Diirig, R. Troger, K. Lilja, K. Wiberg, and L. Ahrens. 2018. Per- and polyfluoroalkyl substances in Swedish groundwater and surface water: Implications for environmental quality standards and drinking water guidelines. Environmental Science & Technology 52:4340-4349. http://dx.doi.org/10.1021/acs.est.7b05718. Labadie, P., and M. Chevreuil. 2011. Biogeochemical dynamics of perfluorinated alkyl acids and sulfonates in the River Seine (Paris, France) under contrasting hydrological conditions. Environmental Pollution 159:3634-3639. http://dx.doi.Org/10.1016/i.envpol.2011.07.028. Lasier, P.J., J.W. Washington, S.M. Hassan, and T.M. Jenkins. 2011. Perfluorinated chemicals in surface waters and sediments from northwest Georgia, USA, and their bioaccumulation in Lumbriculus variegatus. Environmental Toxicology and Chemistry 30:2194-2201. http://dx.doi.org/10.10Q2/etc.622. Lescord, G.L., K.A. Kidd, A.O. De Silva, M. Williamson, C. Spencer, X. Wang, and D.C. Muir. 2015. Perfluorinated and polyfluorinated compounds in lake food webs from the Canadian high Arctic. Environmental Science and Technology 49:2694-2702. http://dx.doi.org/10.1021/es5Q48649. Lindstrom, A.B., M.J. Strynar, A.D. Delinsky, S.F. Nakayama, L. McMillan, E.L. Libelo, M. Neill, and L. Thomas. 2011. Application of WWTP biosolids and resulting perfluorinated compound contamination of surface and well water in Decatur, Alabama, USA. Environmental Science & Technology 45:8015-8021. http://dx.doi.org/10.1021/eslQ39425. Loos, R., S. Tavazzi, G. Mariani, G. Suurkuusk, B. Paracchini, and G. Umlauf. 2017. Analysis of emerging organic contaminants in water, fish and suspended particulate matter (SPM) in the Joint Danube Survey using solid-phase extraction followed by UHPLC-MS-MS and GC-MS analysis. Science of the Total Environment 607-608:1201-1212. http://dx.doi.Org/10.1016/i.scitotenv.2017.07.039. Lorenzo, M., J. Campo, M. Farre, F. Perez, Y. Pico, and D. Barcelo. 2015. Perfluoroalkyl substances in the Ebro and Guadalquivir river basins (Spain). Science of the Total Environment 540:191-199. http://dx.doi.Org/10.1016/i.scitotenv.2015.07.045. Moller, A., L. Ahrens, R. Surm, J. Westerveld, F. van der Wielen, R. Ebinghaus, and P. de Voogt. 2010. Distribution and sources of polyfluoroalkyl substances (PFAS) in the River Rhine watershed. Environmental Pollution 158:3243-3250. http://dx.doi.Org/10.1016/i.envpol.2010.07.019. 64 ------- Munoz, G., L.C. Fechner, E. Geneste, P. Pardon, H. Budzinski, and P. Labadie. 2016. Spatio- temporal dynamics of per and polyfluoroalkyl substances (PFASs) and transfer to periphytic biofilm in an urban river: Case-study on the River Seine. Environmental Science and Pollution Research 25:23574-23582. http://dx.doi.org/10.1007/sll356-016-8Q51-9. Mussabek, D., L. Ahrens, K.M. Persson, and R. Berndtsson. 2019. Temporal trends and sediment-water partitioning of per- and polyfluoroalkyl substances (PFAS) in lake sediment. Chemosphere 227:624-629. http://dx.doi.Org/10.1016/i.chemosphere.2019.04.074. Nakayama, S., M.J. Strynar, L. Helfant, P. Egeghy, X. Ye, and A.B. Lindstrom. 2007. Perfluorinated compounds in the Cape Fear Drainage Basin in North Carolina. Environmental Science & Technology 41:5271-5276. http://dx.doi.org/10.1021/es070792y. Nakayama, S.F., M.J. Strynar, J.L. Reiner, A.D. Delinsky, and A.B. Lindstrom. 2010. Determination of perfluorinated compounds in the Upper Mississippi River Basin. Environmental Science & Technology 44:4103-4109. http://dx.doi.org/10.1021/esl00382z. Newsted, J.L., R. Holem, G. Hohenstein, C. Lange, M. Ellefson, W. Reagen, and S. Wolf. 2017. Spatial and temporal trends of poly- and perfluoroalkyl substances in fish fillets and water collected from pool 2 of the Upper Mississippi River. Environmental Toxicology and Chemistry 36:3138-3147. http://dx.doi.org/10.10Q2/etc.3891. Newton, S., R. McMahen, J.A. Stoeckel, M. Chislock, A. Lindstrom, and M. Strynar. 2017. Novel polyfluorinated compounds identified using high resolution mass spectrometry downstream of manufacturing facilities near Decatur, Alabama. Environmental Science & Technology 51:1544-1552. http://dx.doi.org/10.1021/acs.est.6b05330. Pan, Y., H. Zhang, Q. Cui, N. Sheng, L.W.Y. Yeung, Y. Sun, Y. Guo, and J. Dai. 2018. Worldwide distribution of novel perfluoroether carboxylic and sulfonic acids in surface water. Environmental Science and Technology 52:7621-7629. https://pubs.acs.org/doi/10.1021/acs.est.8b00829. Post, G.B., J.B. Louis, R.L. Lippincott, and N.A. Procopio. 2013. Occurrence of perfluorinated compounds in raw water from New Jersey public drinking water systems. Environmental Science & Technology 47:13266-13275. http://dx.doi.org/10.1021/es4Q2884x. Procopio, N.A., R. Karl, S.M. Goodrow, J. Maggio, J.B. Louis, and T.B. Atherholt. 2017. Occurrence and source identification of perfluoroalkyl acids (PFAAs) in the Metedeconk River Watershed, New Jersey. Environmental Science and Pollution Research 24:27125- 27135. http://dx.doi.org/10.1007/sll356-017-03Q9-3. 65 ------- Rostkowski, P., S. Taniyasu, N. Yamashita, J.J. Falandysz, t. Zegarowski, A. Chojnacka, K. Pazdro, and J. Falandysz. 2009. Survey of perfluorinated compounds (PFCs) in surface waters of Poland. Journal of Environmental Science and Health, Part A, Toxic/Hazardous Substances and Environmental Engineering 44:1518-1527. http://dx.doi.org/10.1080/1093452090326333Q. Shafique, U., S. Schulze, C. Slawik, A. Bohme, A. Paschke, and G. Schuurmann. 2017. Perfluoroalkyl acids in aqueous samples from Germany and Kenya. Environmental Science and Pollution Research 24:11031-11043. http://dx.doi.org/10.1007/sll356-016-7Q76-4. Subedi, B., N. Codru, D.M. Dziewulski, L.R. Wilson, J. Xue, S. Yun, E. Braun-Howland, C. Minihane, and K. Kannan. 2015. A pilot study on the assessment of trace organic contaminants including pharmaceuticals and personal care products from on-site wastewater treatment systems along Skaneateles Lake in New York State, USA. Water Research 72:28-39. https://doi.Org/10.1016/i.watres.2014.10.049. Valsecchi, S., M. Rusconi, M. Mazzoni, G. Viviano, R. Pagnotta, C. Zaghi, G. Serrini, and S. Polesello. 2015. Occurrence and sources of perfluoroalkyl acids in Italian river basins. Chemosphere 129:126-134. http://dx.doi.Org/10.1016/i.chemosphere.2014.07.044. Veillette, J., D.C.G. Muir, D. Antoniades, J.M. Small, C. Spencer, T.N. Loewen, J.A. Babaluk, J.D. Reist, and W.F. Vincent. 2012. Perfluorinated chemicals in meromictic lakes on the northern coast of Ellesmere Island, High Arctic Canada. Arctic 65:245-256. https://doi.org/10.14430/arctic4213. Wagner, A., B. Raue, H.J. Brauch, E. Worch, and F.T. Lange. 2013. Determination of adsorbable organic fluorine from aqueous environmental samples by adsorption to polystyrene- divinylbenzene based activated carbon and combustion ion chromatography. Journal of Chromatography A 1295:82-89. http://dx.doi.Org/10.1016/i.chroma.2013.04.051. Wilkinson, J.L., P.S. Hooda, J. Swinden, J. Barker, and S. Barton. 2017. Spatial distribution of organic contaminants in three rivers of Southern England bound to suspended particulate material and dissolved in water. Science of the Total Environment 593- 594:487-497. http://dx.doi.Org/10.1016/i.scitotenv.2017.03.167. Yeung, L.W.Y., C. Stadey, and S.A. Mabury. 2017. Simultaneous analysis of perfluoroalkyl and polyfluoroalkyl substances including ultrashort-chain C2 and C3 compounds in rain and river water samples by ultra performance convergence chromatography. Journal of Chromatography A 1522:78-85. http://dx.doi.Org/10.1016/i.chroma.2017.09.049. Zhang, X., R. Lohmann, C. Dassuncao, X.C. Hu, A.K. Weber, C.D. Vecitis, and E.M. Sunderland. 2016. Source attribution of poly- and perfluoroalkyl substances (PFASs) in surface waters from Rhode Island and the New York metropolitan area. Environmental Science & Technology Letters 3: 316-321. http://dx.doi.org/10.1021/acs.estlett.6b00255. 66 ------- Zhao, Z., Z. Xie, J. Tang, R. Sturm, Y. Chen, G. Zhang, and R. Ebinghaus. 2015. Seasonal variations and spatial distributions of perfluoroalkyl substances in the rivers Elbe and lower Weser and the North Sea. Chemosphere 129:118-125. http://dx.doi.Org/10.1016/i.chemosphere.2014.03.050. 67 ------- Appendix B: Bioaccumulation Factor (BAF) Supporting Information Search Strings used for literature review of PFBS bioaccumulation data: ("375-73-5" OR "29420-49-3" OR "45187-15-3" OR "Perfluorobutane Sulfonic Acid" OR PFBS OR "Potassium Perfluorobutane Sulfonate" OR "K+PFBS" OR "nonafluorobutane-l-sulfonic acid" OR "1,1,2,2,3,3,4,4,4-Nonafluoro-l-butanesulfonic acid" OR "Nonafluorobutane sulfonic Acid" OR "Perfluro-l-butanesulfonate" OR "1-Perfluorobutanesulfonic acid" OR "Nonafluoro-l-butanesulfonic acid" OR "Nonafluorobutanesulfonyl fluoride" OR "Nonafluorobutane-l-sulfonatato" OR "Nonafluorobutanesulfonic acid" OR "Perfluorobutanesulfonic acid" OR "Perfluorobutane sulfonic acid" OR "1,1,2,2,3,3,4,4,4- Nonafluorobutane-l-sulphonic acid" OR "1,1,2,2,3,3,4,4,4-Nonafluoro-l-butanesulfonyl fluoride" OR "1,1,2,2,3,3,4,4,4-nonafluorobutane-l-sulfonyl fluoride" OR "Perfluorobutanesulfonate" OR "Perfluorobutane Sulfonate" OR "Perfluorobutanesulfonyl fluoride" OR "1-Butanesulfonic acid, 1,1,2,2,3,3,4,4,4-nonafluoro-" OR "1-butanesulfonic acid, 1,1,2,2,3,3,4,4,4-nonafluror, potassium salt" OR "1-Butanesulfonic acid, nonafluoro-" OR "Perfluoro-l-butanesulfonate" OR "Perfluorobutylsulfonate" OR "Perfluoro-1- butanesulfonyl fluoride" OR "Potassium;l,l,2,2,3,3,4,4,4-nonafluorobutane-l-sulfonate" OR "Potassium nonafluoro-l-butanesulfonate" OR "Ammonium nonafluorobutane-l-sulfonate" OR "Ammonium perfluorobutanesulfonate" OR "Potassium perfluorobutanesulfonate" OR "Potassium perfluorobutane sulfonate" OR "Potassium nonafluorobutane-l-sulfonate" OR "Potassium nonafluoro-l-butanesulfonate" OR "Potassium PFBS" OR "PFBuS" OR "C539348" OR "1FV02N6NVO" OR "DTXSID5030030" OR "FC-98" OR "EFTOP FBSA" OR "UNII-1FV02N6NVO" OR "SCHEMBL23932" OR "CHEMBL1198521" OR "HSDB 8294" OR "CHEBI:132446" OR "CS-B0899" OR "MFCD01320794" OR "AKOS015852768" OR "NCI60_006096" OR "FT-0676348" OR "FT-0676859" OR "N0709" OR "D77221" OR "Q410426") AND ("Bioaccumulation Factor" OR "Bioconcentration Factor" OR bcf OR baf OR bioaccumulation OR bioconcentration OR uptake OR depuration OR accumulation) 68 ------- BAF Calculation Description for PFBS The EPA used the decision framework presented in the Methodology for Deriving Ambient Water Quality Criteria for the Protection of Human Health (2000), Technical Support Document, Volume 2: Development of National Bioaccumulation Factors (Technical Support Document, Volume 2) (EPA, 2003) to identify procedures to derive national trophic level-specific BAFs for PFBS based on that chemical's properties (e.g., ionization, hydrophobicity), metabolism, and biomagnification potential (see Figure 1). The EPA followed the guidelines provided in Section 5.5 of EPA's Methodology for Deriving Ambient Water Quality Criteria for the Protection of Human Health (2000) (EPA's 2000 Methodology) (EPA, 2000), to assess the occurrence of cationic and anionic forms of PFBS at typical environmental pH ranges. PFBS is a nonionic organic chemical (with ionization significant at typical environmental pH ranges) (EPA, 2021a,b). As explained in Section 5.5 of EPA's 2000 Methodology (EPA, 2000), when a significant fraction of the total chemical concentration is expected to be present as the ionized species in water, procedures for deriving the national BAF rely on empirical (measured) methods (i.e., Procedures 5 and 6). The EPA followed the guidelines in Sections 3.2.1 and 3.2.2 of the Technical Support Document, Volume 2, to evaluate the biomagnification potential of PFBS. Based on the information in Loi et al. (2011), it was determined that biomagnification was unlikely. Based on the characteristics of PFBS, the EPA selected Procedure 5 for deriving national BAF values for this chemical. As described in Section 4.2.1, for a given procedure, the EPA selected the method that provided BAF estimates for all three TLs (TL 2-TL 4) in the following priority: • BAF estimates using the BAF method (i.e., based on field-measured BAFs) if possible. • BAF estimates using the BCF method if (a) the BAF method did not produce estimates for all three TLs and (b) the BCF method produced national-level BAF estimates for all three TLs. The EPA was able to locate field-measured BAFs for TLs 2, 3, and 4 for PFBS from the peer- reviewed literature sources for which sufficient information was provided to determine the quality and usability of the data. Therefore, the EPA used the BAF method (EPA, 2003) to derive the national BAF values for this chemical. Calculating Baseline BAFs As described in Section 4.2.3, the national-level BAF equation adjusts the TL baseline BAFs for nonionic organic chemicals by national default values for lipid content, as well as dissolved and particulate organic carbon content. However, the partitioning of PFBS is related to protein binding properties (ATSDR, 2021; ECHA, 2019). The EPA considered protein-normalizing the measured BAF values in the baseline BAF equation; however, insufficient data were available from the scientific literature on protein content of aquatic organisms and on the binding efficiencies of PFBS to various proteins in aquatic organisms. Because of this lack of data on the relationship between protein content and PFBS bioaccumulation, attempts to normalize BAFs based on protein content would likely introduce greater uncertainty into BAF averages. 69 ------- Consistent with the EPA's 2000 Methodology (EPA, 2000), a procedure analogous to the one used to adjust for the water-dissolved portions of a nonionic organic chemical is applied to the measured BAFs for PFBS. As described in EPA's (2003) Technical Support Document, Volume 2, the Kpoc (the equilibrium partition coefficient of the chemical between the particulate organic carbon [POC] phase and the freely dissolved phase of water) is approximately equal to the Kow of a hydrophobic organic chemical. It is further described in the EPA's (2003) Technical Support Document, Volume 2, that Kdoc (the equilibrium partition coefficient of the chemical between the dissolved organic carbon [DOC] phase and the freely dissolved phase of water) is directly proportional to the Kow of a hydrophobic organic chemical, and that Kdoc is less than the Kow. The log Koc for PFBS provided in ATSDR (2021) is 2.06 (as determined from a groundwater aquifer study) and was used in the national BAF calculations to adjust for the water-dissolved portions of a nonionic organic chemical. The EPA determined that the Koc values were applicable to POC but there is no indication that they would be applicable to DOC. Thus, the amount of PFBS partitioned to DOC was presumed to be part of the aqueous fraction of the ffd equation, resulting in the following formula (Eq. 1): ffd_ (Eq. 1) n+ (POCKoc)l Where: • ffd = fraction of the total concentration of chemical in water that is freely dissolved. • POC = national default value of 0.5 mg/L (refer to page 5-44 of the EPA's 2000 Methodology [EPA, 2000]) is used in baseline BAF calculations, unless this value is reported in the BAF source. • Koc = PFBS log Koc; log koc = 2.06 (ATSDR, 2021). Because the measured BAFs for PFBS are not adjusted for lipid or protein content, the baseline BAF equation (refer to Eq. 5-10 on pages 5-24 and 5-25 of the EPA's 2000 Methodology [EPA, 2000]) is adjusted (as shown below in Eq. 2) to determine the freely dissolved PFBS in water: .. _ . _ Measured BAF . Baseline BAF = 1 (Eq. 2) ffd The EPA used this equation to calculate baseline BAFs from field measured BAFs based on total concentrations. Dissolved PFBS Baseline BAFs The EPA included results from several field BAF studies for PFBS reported as dissolved (i.e., filtered) concentrations in its baseline BAF calculations. Because these dissolved PFBS data are presumed to represent the freely-dissolved (non-particulate) fraction, the ffd term in Eq. 2 is set to 1. Also, as described above, the measured BAFs for PFBS are not being adjusted for lipid or protein content to calculate baseline BAFs for PFBS. Thus, Eq. 3 is used to calculate the freely dissolved concentration of PFBS for "baseline BAFs" using field-measured dissolved PFBS BAFs: Baseline BAF = Measured (dissolved) BAF — 1 (Eq. 3) 70 ------- Calculating the National BAFs Final baseline BAFs were used to compute national BAFs for PFBS. Eq. 4 (an equation analogous to the equation used for nonionic organic chemicals in the EPA's 2015 Updated Human Health criteria for calculating national BAFs (see Eq. 5-28 on Page 5-42 of the EPA's 2000 Methodology [EPA, 2000]) is used to convert the baseline BAF to a national BAF for each trophic level: National BAF(XLn) = [(Final Baseline BAFfd)TLn + 1] • (ffd) (Eq. 4) Where: • National BAF = national BAF (L/kg-tissue). • (Final Baseline BAF)-n_n = mean baseline BAF forTL "n" (L/kg-lipid). • ffd = fraction of the total concentration of chemical in water that is freely dissolved. In summary, for PFBS, the baseline BAFs are calculated using Equation 2 (for field measured BAFs calculated from total water concentrations) and Equation. 3 (for field BAFs calculated from dissolved water concentrations) for each TL. National BAFs are then calculated from TL baseline BAFs using Equation 4 as shown below. National Trophic level BAF calculations: National BAF PFBS(TL2) = [(355.9)TL2 + 1] x (0.9999) = 356.9 L/kg = 360 L/kg (rounded) National BAF PFBS(TL3) = [(285.6)TL3 + 1] x (0.9999) = 286.6 L/kg = 290 L/kg (rounded) National BAF PFBS(TL4) = [(866.9)TL4 + 1] x (0.9999) = 867.8 L/kg = 870 L/kg (rounded) References ATSDR (Agency for Toxic Substances and Disease Registry). 2021. Toxicological Profile for Perfluoroalkyls. U.S. Department of Health and Human Services, ATSDR, Atlanta, GA. Accessed January 2024. https://www.atsdr.cdc.gov/toxprofiles/tp200.pdf. ECHA (European Chemicals Agency). 2019. Support Document for Identification of Perfluorobutane Sulfonic Acid and its Salts as Substances of Very High Concern Because of Their Hazardous Properties Which Cause Probable Serious Effects to Human Health and the Environment Which Give Rise to An Equivalent Level of Concern to Those ofCMR and PBT/vPvB Substances (Article 57F). Accessed February 2024. https://echa.europa.eu/documents/10162/891ab33d-d263-cc4b-0f2d-d84cfb7f424a. 71 ------- EPA (Environmental Protection Agency). 2000. Methodology for Deriving Ambient Water Quality Criteria for the Protection of Human Health (2000). EPA-822-B-00-004. EPA, Office of Water, Office of Science and Technology, Washington, DC. Accessed January 2024. https://www.epa.gov/sites/default/files/2018-10/documents/methodology-wqc- protection-hh-2000.pdf. EPA (Environmental Protection Agency). 2003. Methodology for Deriving Ambient Water Quality Criteria for the Protection of Human Health (2000), Technical Support Document Volume 2: Development of National Bioaccumulation Factors. EPA-822-R-03-030. EPA, Office of Water, Office of Science and Technology, Washington, DC. Accessed January 2024. https://www.epa.gov/sites/default/files/2018-10/documents/methodology-wqc- protection-hh-2000-volume2.pdf. EPA (Environmental Protection Agency). 2021a. Provisional Peer-Reviewed Toxicity Values for Perfluorobutane Sulfonic Acid (PFBS) and Related Compound Potassium Perfluorobutane Sulfonate. EPA/690/R-21/001. EPA, Office of Research and Development, Center for Public Health and Environmental Assessment, Cincinnati, OH. Accessed February 2024. https://cfpub.epa.gov/ncea/pprtv/recordisplay.cfm?deid=350061. EPA (Environmental Protection Agency). 2021b. Human Health Toxicity Values for Perfluorobutane Sulfonic Acid (CASRN 375-73-5) and Related Compound Potassium Perfluorobutane Sulfonate (CASRN 29420-49-3). EPA-600-R-20-345F. EPA, Office of Research and Development, Washington, DC. Accessed February 2024. https://www.epa.gov/pfas/learn-about-human-health-toxicity-assessment-pfbs. Loi, E.I.H., L.W.Y. Yeung, S. Taniyasu, P.K.S. Lam, K. Kannan, and N. Yamashita. 2011. Trophic magnification of poly- and perfluorinated compounds in a subtropical food web. Environmental Science & Technology 45:5506-5513. https://pubmed.ncbi.nlm.nih.gov/21644538/. 72 ------- Appendix C: Supporting Literature for Deriving the Relative Source Contribution Table C-l. Compilation of studies describing PFBS occurrence in food. Study Location and Source Food Types Results North America Blaine et al. United States (Midwestern) Fruits and ND in corn, lettuce, tomato in (2013) Greenhouse and field studies, unamended controls vegetables, grain unamended soil Blaine et al. United States (Midwestern) Fruits and Radish root: DF NR, (2014) Greenhouse study, unamended controls vegetables mean = 22.36 ng/g ND in celery shoot, pea fruit Byrne et al. United States (Alaska) Seafood Blackfish: DF 48%, range = ND- (2017) Upstream/downstream of former defense site (Suqi River) 59.2 ng/g ww Highest concentration was upstream Schecter et al. United States (Texas) Dairy, fruits and Cod: DF NR, mean = 0.12 ng/g ww (2010) Grocery stores vegetables, grain, meat, seafoodh, fats/other ND in salmon, canned sardines, canned tuna, fresh catfish fillet, frozen fish sticks, tilapia, cheeses (American, mozzarella, Colby, cheddar, Swiss, provolone, and Monterey jack), butter, cream cheese, frozen yogurt, ice cream, whole milk, whole milk yogurt, potatoes, apples, cereals, bacon, canned chili, ham, hamburger, roast beef, sausages, sliced chicken breast, sliced turkey, canola oil, margarine, olive oil, peanut butter, eggs Scher etal. United States (Minnesota) Fruits and Within GCA: (2018) Home gardens Near former 3M PFAS production facility, homes within and outside a GCA vegetables Leaf: DF 6%, max = 0.061 ng/g Stem: DF 4%, max = 0.065 ng/g ND in floret, fruit, root, seed Outside GCA: ND Young et al. United States (17 states) Dairy ND in retail cow's milk (2012) Retail markets Young et al. United States (Maryland, Seafood ND in crab, shrimp, striped bass, (2013) Mississippi, Tennessee, Florida, New York, Texas, Washington, D.C.) Retail markets farm raised catfish, farm raised salmon h Some PFBS dietary studies use the term "seafood" to indicate fish and shellfish from ocean, freshwater, or estuarine water bodies. Information about the water bodies assessed in individual studies is reported in the articles. 73 ------- Study Location and Source Food Types Results Europe Barbosa et al. Belgium, France, the Seafood ND in raw and steamed fish (P. (2018) Netherlands, Portugal Various markets platessa, M. australis, M. capenis, K. pelamis, and M. edulis) D'Hollander et Belgium, Czech Republic, Fruit, cereals, Sweets: DFa25%, range = ND- al. (2015) Italy, Norway PERFOOD study; items from 3 national retail stores of different brands and countries of origin sweets, salt 0.0016 ng/g Fruit: DFa 19%, range = ND- 0.067 ng/g ND in cereals, salt Domingo et al. Spain (Catalonia) 12 food Vegetables: DF NR, (2012) Local markets, small stores, supermarkets, big grocery stores categories mean = 0.013 ng/g fw Fish and seafood: DF NR, mean = 0.054 ng/g fw ND in meat and meat products, tubers, fruits, eggs, milk, dairy products, cereals, pulses, industrial bakery, oils Ericson et al. Spain 18 food ND in all categories: veal, pork, (2008) Local markets, large supermarkets, grocery stores categories chicken, lamb, white fish, seafood, tinned fish, blue fish, whole milk, semi-skimmed milk, dairy products, vegetables, pulses, cereals, fruits, oil, margarine, and eggs Eriksson et al. Denmark Dairy, fruits and Milk: (2013) Farm, dairy farm, fish from vegetables, Farmer (Havnardal): Faroe Shelf area seafood point = 0.019 ng/g ww Dairy (Faroe Island): point = 0.017 ng/g ww; ND or NQ in 4 samples ND in yogurt, creme fraiche, potatoes, farmed salmon, wild- caught cod, wild-caught saithe Eschauzier et al. The Netherlands Fats/other Brewed coffee (manual): mean (2013) (Amsterdam) Cafes, universities, supermarkets (range) = 1.6 (1.3-2.0) ng/L Brewed coffee (machine): mean (range) = 2.9 (ND-9.8) ng/L Cola: mean (range) = 7.9 (ND- 12) ng/L Falandysz et al. Poland Meat, seafood ND in eider duck, cod (2006) Gulf of Gdansk, Baltic Sea south coast 74 ------- Study Location and Source Food Types Results Gebbink et al. Sweden 12 food ND in all categories: dairy products, (2015) Major grocery chain stores, market basket samples categories meat products, fats, pastries, fish products, egg, cereal products, vegetables, fruit, potatoes, sugar and sweets, soft drinks Herzke et al. Belgium, Czech Republic, Vegetables ND for all vegetables (2013) Italy, Norway PERFOOD study: items from 3 national retail stores of different brands per location Hlouskova et al. Belgium, Czech Republic, Pooled DF 5%, mean (range) = 0.00975 (2013) Italy, Norway Several national supermarkets milk/dairy products, meat, fish, hen eggs (0.006-0.012) ng/g Holzer etal. Germany Seafood Lake Mohne /River Mohne: ND in (2011) Fish from Lake Mohne and river Mohne, contaminated with PFCs from use of polluted soil conditioner on agricultural lands; retail trade, wholesale trade, supermarkets, and producers cisco, eel, perch, pike, and roach Trade/markets: ND in eel, pike/perch, and trout Jogsten et al. Spain (Catalonia) Fruits and ND in lettuce, raw, cooked, and (2009) Local markets, large vegetables, fried meat (veal, pork, and chicken), supermarkets, grocery meat, seafood, fried chicken nuggets, black stores fats/other pudding, lamb liver, pate of pork liver, foie gras of duck, "Frankfurt" sausages, home-made marinated salmon, and common salt Jorundsdottir et Iceland Seafood ND in anglerfish, Atlantic cod, blue al. (2014) Collected during biannual scientific surveys, commercially produced whiting, lemon sole, ling, lumpfish, plaice, and pollock Lankova et al. Czech Republic Fats/other ND in infant formula (2013) Retail market Noorlander et The Netherlands 15 food ND in all categories: flour, fatty fish, al. (2011) Several Dutch retail store chains with nationwide coverage categories lean fish, pork, eggs, crustaceans, bakery products, vegetables/fruit, cheese, beef, chicken/poultry, butter, milk, vegetable oil, and industrial oil 75 ------- Study Location and Source Food Types Results Papadopoulou Norway Solid foods Solid foods (unspecific food et al. (2017) A-TEAM project: food and (11 food category): DF 2%, range = ND- drinks collected by categories), 0.001 ng/g participants as duplicate liquid foods ND in liquid foods (coffee, tea and diet samples (5 drinks) cocoa, milk, water, alcoholic beverages and soft drinks) Perez et al. Serbia (Belgrade and Novi 8 food Categories included cereals, pulses (2014) Sad), Spain (Barcelona, Girona, and Madrid) Various supermarkets and retail stores categories and starchy roots, tree-nuts, oil crops and vegetable oils, vegetables and fruits, meat and meat products, milk, animal fats, dairy products, and eggs, fish and seafood, and others such as candies or coffee Spain: DF 3.2%, range = ND- 13 ng/g (primarily fish, oils) Serbia: DF 5.2%, range = ND- 0.460 ng/g (primarily meat and meat products, cereals) Riviere et al. France Seafood, ND in infant food, vegetables, (2019) Based on results of national consumption survey fats/other nonalcoholic beverages, dairy- based desserts, milk, mixed dishes, fish, ultra-fresh dairy products, meat, poultry and game Scordo et al. Italy Fruits Olives: DFa 100%, mean3 (2020) Supermarkets (range) = 0.294 (0.185- 0.403) ng/g dw ND in strawberries Surma et al. Spain, Slovakia Fats/other Spices: ND-1.01 ng/g (2017) Source NR Spain: Detected in anise, star anise, fennel, coriander, cinnamon, peppermint, parsley, thyme, laurel, cumin, and oregano ND in white pepper, cardamon, clove, nutmeg, allspice, vanilla, ginger, garlic, black paper, and hot pepper (mild and hot) Slovakia: ND in anise, star anise, white pepper, fennel, cardamom, clove, coriander, nutmeg, allspice, cinnamon, vanilla, and ginger Sznajder- Poland Fruits and ND in apples, bananas, cherries, Katarzyriska et Markets vegetables lemons, oranges, strawberries, al. (2018) beetroots, carrots, tomatoes, potatoes, and white cabbage 76 ------- Study Location and Source Food Types Results Sznajder- Poland Dairy All dairy: sum PFBS = 0.04 ng/g Katarzyriska et Markets Butter: range = 0.01-0.02 ng/g al. (2019) ND in camembert-type cheese, cottage cheese, milk, natural yogurt, sour cream, kefir (bonny clabber) Vassiliadou et al. Greece Seafood Hake: raw mean = 0.45 ng/g ww, (2015) Local fish markets, mariculture farm, fishing sites fried mean = 0.83 ng/g ww Shrimp: raw mean = 1.37 ng/g ww ND in raw, fried, and grilled anchovy, bogue, picarel, sand smelt, sardine, squid, striped mullet, raw and fried mussel, fried shrimp, and grilled hake Zafeiraki et al. Greece, the Netherlands Fats/other ND in chicken eggs (2016a) Home and commercially produced Zafeiraki et al. The Netherlands Meat ND for horse, sheep, cow, pig, and (2016b) Local markets and slaughterhouses chicken liver Multiple Continents Chiesa et al. United States (Pacific Seafood ND in wild-caught salmon (2019) Ocean) Wholesale fish market Canada Seafood ND in wild-caught salmon Wholesale fish market Norway Seafood ND in farm salmon Wholesale fish market Scotland Seafood ND in wild-caught and farm salmon Wholesale fish market Notes: DF = detection frequency; dw = dry weight; fw = fresh weight; GCA = groundwater contamination area; ND = not detected; ng/g = nanogram per gram; ng/L = nanogram per liter; NR = not reported; PFAS = per- and polyfluoroalkyl substances; NQ = not quantified; ng/L = microgram per liter; ww = wet weight. Bold indicates detected levels of PFBS in food. a The DF and/or mean was not reported in the study and was calculated in this synthesis. Means were calculated only when DF = 100%. 77 ------- Table C-2. Compilation of studies describing PFBS occurrence in indoor dust. Study Location Site Details Results North America Byrne et al. (2017) United States (St. Lawrence Island, Alaska) Homes (49) DF 16%, median = ND; 95th percentile = 1.76 ng/g Fraser et al. United States Homes (30); offices (31); Homes: DF 3% (single (2013) (Boston, Massachusetts) vehicles (13) detection), range = ND- 4.98 ng/g Offices: DF 10%, range = ND- 12.0 ng/g Vehicles: DF 0% Knobeloch et al. United States (Great Homes (39) DF 59%, median (range) = 1.8 (2012) Lakes Basin, Wisconsin) (ND-31) ng/g Kubwabo et al. Canada (Ottawa) Homes (67) DF 0% (2005) Scher et al. (2019) United States (Twin Near former 3M PFAS Entryway: DF 11%, median Cities metropolitan production facility; (range) = ND (ND-58 ng/g) region, Minnesota) 19 homes within the GCA Living room: DF 16%, median (range) = ND (ND-58 ng/g) Strynarand United States (Cities Homes (102) and DF 33%, mean Lindstrom (2008) in North Carolina and Ohio) daycare centers (10); samples had been collected in 2000-2001 during EPA's Children's Total Exposure to Persistent Pesticides and Other Persistent Organic Pollutants (CTEPP) study (range) = 41.7 (ND-1,150) ng/g Zheng et al. (2020) United States Childcare facilities DF 90%, mean (range) = 0.34 (Seattle, Washington (20 samples from (ND-0.86) ng/g and West Lafayette, 7 facilities in Seattle and Indiana) 1 in West Lafayette) Europe de la Torre et al. Spain (unspecified), Homes (65) Spain: DF 52%, median (2019) Belgium (unspecified), Italy (unspecified) (range) = 0.70 (ND-12.0) ng/g Belgium: DF 27%, median (range) = 0.40 (ND-56.7) ng/g Italy: DF 18%, median (range) = 0.40 (ND-11.6) ng/g D'Hollander et al. Belgium (Flanders) Homes (45); offices (10) Homes: DF 47%, (2010) median = 0 ng/g dw Offices: DF NR, median = 0.2 ng/g dw Giovanoulis et al. Sweden (Stockholm) Preschools (20) DF 0% (2019) 78 ------- Study Location Site Details Results Harrad et al. Ireland (Dublin, Homes (32); offices (33); Homes: DF 81%, mean (2019) Galway, and Limerick cars (31); (range) = 17 (ND-110) ng/g counties) classrooms (32) Offices: DF 88%, mean (range) = 19 (ND-98) ng/g Cars: DF 75%, mean (range) = 12 (ND-170) ng/g Classrooms: DF 97%, mean (range) = 17 (ND-49) ng/g Haug et al. (2011) Norway (Oslo) Homes (41) DF 22%, mean (range) = 1.3 (0.17-9.8) ng/g Huber et al. (2011) Norway (Troms0) Homes (7; carpet, bedroom, sofa); one office; one storage room that had been used for storage of "highly contaminated PFC [polyfluorinated compounds] samples and technical mixtures for several years" All homes: DF NR, median = 1.1 ng/g Living room: DFa 57%, range = ND-10.6 ng/g Carpet, bedroom, sofa: DF 0% Office: point = 3.8 ng/g Storage room: point = 1,089 ng/g Jogsten et al. Spain (Catalonia) Homes (10) DF 60%, range = ND-6.5 ng/g (2012) Padilla-Sanchez Norway (Oslo) Homes (7) DF 14% (single detection), and Haug(2016) range = ND-3 ng/g Winkens et al. Finland (Kuopio) Homes (63 children's DF 12.7%, median (range) = ND (2018) bedrooms) (ND-13.5) ng/g Multiple Continents Karaskova et al. United States Homes (14) DF 60%, mean (range) = 1.4 (2016) (unspecified) (ND-2.6) ng/g Canada (unspecified) Homes (15) DF 55%, mean (range) = 1.6 (ND-5.8) ng/g Czech Republic Homes (12) DF 37.5%, mean (unspecified) (range) = 3.6 (ND-14.4) ng/g Kato et al. (2009) United States (Atlanta, Georgia), Germany (unspecified), United Kingdom (unspecified), Australia (unspecified) Homes (39) DF 92.3%, median (range) = 359 (ND-7,718) ng/g Notes: DF = detection frequency; GCA = groundwater contamination area; ND = not detected; ng/g = nanogram per gram; NR = not reported; dw = dry weight. aThe DF and/or mean was not reported in the study and was calculated in this synthesis. Means were calculated only when DF = 100%. 79 ------- Table C-3. Compilation of studies describing PFBS occurrence in soil. Study Location Site Details Results North America Anderson et al. United States Ten U.S. Air Force Surface soil: DF 35%, median (2016) (unspecified) installations with historic AFFF release, surface and subsurface soils (range) = 0.775 (ND-52.0) ng/g Subsurface soil: DF 35%, median (range) = 1.30 (ND-79.0) ng/g Blaine et al. (2013) United States Urban and rural full-scale Urban control: DF NR, (Midwestern) field study control (nonamended) soil mean = 0.10 ng/g Rural control: DF NR, mean = ND Cabrerizo et al. Canada (Melville Catchment areas of lakes DF 100%, mean3 (2018) and Cornwallis Islands) (range) = 0.0024 (0.0004- 0.0083) ng/g dw Dreyer et al. Canada (Ottawa, Mer Bleue Bog Peat Detected once at 0.071 ng/g in (2012) Ontario) samples (core samples) 1973 sample and not considered for further evaluation Eberle et al. (2017) United States (Joint Firefighting training site, Pretreatment: DF 60%, Base Langley-Eustis, pre- and posttreatment range = 0.61-6.4 ng/g Virginia) Posttreatment: DF 100%, range = 0.07-0.83 ng/g Mejia-Avendano Canada (Lac- Site of 2013 Lac- Background: DF NR, et al. (2017) Megantic, Quebec) Megantic train accident (oil and AFFF runoff area [sampled 2013], burn site and adjacent area [sampled 2015]) mean = 0.035 ng/g dw 2013: DF 75%, mean range = ND-3.15 ng/g dw 2015: DF 36%, mean range = ND-1.25 ng/g dw Nickerson et al. United States Two AFFF-impacted soil Core E: DFa 91%, range = ND- (2020) (unspecified) cores from former fire- training areas 27.37 ng/g dw Core F: DF 100%, range = 0.13- 58.44 ng/g dw Scher et al. (2018) United States (Twin Near former 3M PFAS Within GCA: DF 9%, median Cities metropolitan production facility, (range) = ND (ND-0.17 ng/g) region, Minnesota) homes within and outside a GCA Outside GCA: DF 17%, median (range) = ND (ND-0.031 ng/g) Scher et al. (2019) United States (Twin Near former 3M PFAS DF 10%, median (p90) = ND Cities metropolitan production facility, (0.02) ng/g region, Minnesota) homes within a GCA Venkatesan and United States Control (nonamended) DF 0% Halden (2014) (Baltimore, Maryland) soil from Beltsville Agricultural Research Center 80 ------- Study Location Site Details Results Europe Dauchy etal. (2019) France (unspecified) Firefighting training site, samples collected in 6 areas collected up to 15-m depth; in areas 2 and 6, foams used more intensely and/or before concrete slab was built Areas 1, 3, 4, and 5 combined: DFa 0-10%, range = ND- 7 ng/g dw, across all depths Area 2: DFa 35%, range = ND- 82 ng/g dw, across all depths Area 6: DFa 55%, range = ND- 101 ng/g dw, across all depths Groffen et al. (2019) Belgium (Antwerp) 3M perfluorochemical plant and 4 sites with increasing distance from plant Plant: DF 92%, mean (range) = 7.84 (ND-33) ng/g dw Vlietbos (1 km from plant): DF 90%, mean (range) = 2.79 (ND-7.04) ng/g dw 2.3 km, 3 km, 11 km from plant: DF 0% Gr0nnestad et al. (2019) Norway (Granasen, Jonsvatnet) Granasen (skiing area); Jonsvatnet (reference site) Skiing area: DF 0%b Reference area: DF 70%, mean (range) = 0.0093 (ND- 0.0385 ng/g dw) Harrad et al. (2020) Ireland (multiple cities) 10 landfills, samples collected upwind and downwind Downwind: DF NR, mean (range) = 0.0059 (ND- 0.044) ng/g dw Upwind: DF NR, mean (range) = 0.0011 (ND- 0.0029) ng/g dw Hale et al. (2017) Norway (Gardermoen) Firefighting training site DF 0% Skaar et al. (2019) Norway (Ny- Alesund) Research facility near firefighting training site Background: DF 0% Contaminated: DF 100%, mean3 (range) = 4.9 (2.64- 7.13) ng/g dw Notes: AFFF = aqueous film-forming foam; DF = detection frequency; dw = dry weight; GCA = groundwater contamination area; km = kilometer; ND = not detected; ng/g = nanogram per gram; NR = not reported; PFAS = per- and polyfluoroalkyl substances; p90 = 90th percentile. aThe DF and/or mean was not reported in the study and was calculated in this synthesis. Means were calculated only when DF = 100%. b Gr0nnestad et al. (2019) reported a DF = 10% but a range, mean, and standard deviation of < LOQ. 81 ------- References Anderson, R.H., G.C. Long, R.C. Porter, and J.K. Anderson. 2016. Occurrence of select perfluoroalkyl substances at U.S. Air Force aqueous film-forming foam release sites other than fire-training areas: Field-validation of critical fate and transport properties. Chemosphere 150:678-685. http://dx.doi.Org/10.1016/i.chemosphere.2016.01.014. Barbosa, V, A.L. Maulvault, R.N. Alves, C. Kwadijk, M. Kotterman, A. Tediosi, M. Fernandez- Tejedor, J.J. Sloth, K. Granby, R.R. Rasmussen, J. Robbens, B. De Witte, L. Trabalon, J.O. Fernandes, S.C. Cunha, and A. Marques. 2018. Effects of steaming on contaminants of emerging concern levels in seafood. Food and Chemical Toxicology 118:490-504. http://dx.doi.Org/10.1016/i.fct.2018.05.047. Blaine, A.C., C.D. Rich, L.S. Hundal, C. Lau, M.A. Mills, K.M. Harris, and C.P. Higgins. 2013. Uptake of perfluoroalkyl acids into edible crops via land applied biosolids: Field and greenhouse studies. Environmental Science & Technology 47:14062-14069. http://dx.doi.org/10.1021/es403094q. Blaine, A.C., C.D. Rich, E.M. Sedlacko, L.S. Hundal, K. Kumar, C. Lau, M.A. Mills, K.M. Harris, and C.P. Higgins. 2014. Perfluoroalkyl acid distribution in various plant compartments of edible crops grown in biosolids-amended soils. Environmental Science & Technology 48:7858-7865. http://dx.doi.org/10.1021/es500Q16s. Byrne, S., S. Seguinot-Medina, P. Miller, V. Waghiyi, F.A. von Hippel, C.L. Buck, and D.O. Carpenter. 2017. Exposure to polybrominated diphenyl ethers and perfluoroalkyl substances in a remote population of Alaska Natives. Environmental Pollution 231:387- 395. http://dx.doi.Org/10.1016/i.envpol.2017.08.020. Cabrerizo, A., D.C.G. Muir, A.O. De Silva, X. Wang, S.F. Lamoureux, and M.J. Lafreniere. 2018. Legacy and emerging persistent organic pollutants (POPs) in terrestrial compartments in the high Arctic: Sorption and secondary sources. Environmental Science & Technology 52:14187-14197. http://dx.doi.org/10.1021/acs.est.8b05011. Chiesa, L.M., M. Nobile, F. Ceriani, R. Malandra, F. Arioli, and S. Panseri. 2019. Risk characterisation from the presence of environmental contaminants and antibiotic residues in wild and farmed salmon from different FAO zones. Food Additives and Contaminants: Part A 36:152-162. http://dx.doi.org/10.1080/19440049.2Q18.1563723. Dauchy, X., V. Boiteux, A. Colin, J. Hemard, C. Bach, C. Rosin, and J.F. Munoz. 2019. Deep seepage of per- and polyfluoroalkyl substances through the soil of a firefighter training site and subsequent groundwater contamination. Chemosphere 214:729-737. https://www.sciencedirect.eom/science/a rticle/abs/pii/S0045653518318514. 82 ------- de la Torre, A., I. Navarro, P. Sanz, and M.L.A. Martinez. 2019. Occurrence and human exposure assessment of perfluorinated substances in house dust from three European countries. Science of the Total Environment 685:308-314. http://dx.doi.Org/10.1016/i.scitotenv.2019.05.463. D'Hollander, W., L. Roosens, A. Covaci, C. Cornelis; H. Reynders, K. van Campenhout, P. de Voogt, and L. Bervoets. 2010. Brominated flame retardants and perfluorinated compounds in indoor dust from homes and offices in Flanders, Belgium. Chemosphere 81:478-487. http://dx.doi.Org/10.1016/i.chemosphere.2010.07.043. D'Hollander, W., D. Herzke, S. Huber, J. Hajslova, J. Pulkrabova, G. Brambilla, S.P. De Filippis, L. Bervoets, and P. de Voogt. 2015. Occurrence of perfluorinated alkylated substances in cereals, salt, sweets and fruit items collected in four European countries. Chemosphere 129:179-185. http://dx.doi.Org/10.1016/i.chemosphere.2014.10.011. Domingo, J.L., I.E. Jogsten, U. Eriksson, I. Martorell, G. Perello, M. Nadal, and B. van Bavel. 2012. Human dietary exposure to perfluoroalkyl substances in Catalonia, Spain. Food Chemistry 135:1575-1582. http://dx.doi.Org/10.1016/i.foodchem.2012.06.054. Dreyer, A., S. Thuens, T. Kirchgeorg, and M. Radk. 2012. Ombrotrophic peat bogs are not suited as natural archives to investigate the historical atmospheric deposition of perfluoroalkyl substances. Environmental Science & Technology 46:7512-7519. http://dx.doi.org/10.1021/es204175y. Eberle, D., R. Ball, and T.B. Boving. 2017. Impact of ISCO treatment on PFAA co-contaminants at a former fire training area. Environmental Science & Technology 51:5127-5136. http://dx.doi.org/10.1021/acs.est.6b06591. Ericson, I., R. Marti-Cid, M. Nadal, B. Van Bavel, G. Lindstrom, and J.L. Domingo. 2008. Human exposure to perfluorinated chemicals through the diet: Intake of perfluorinated compounds in foods from the Catalan (Spain) Market. Journal of Agricultural and Food Chemistry 56:1787-1794. http://dx.doi.org/10.1021/if0732408. Eriksson, U., A. Karrman, A. Rotander, B. Mikkelsen, and M. Dam. 2013. Perfluoroalkyl substances (PFASs) in food and water from Faroe Islands. Environmental Science and Pollution Research 20:7940-7948. http://dx.doi.org/10.1007/sll356-013-170Q-3. Eschauzier, C., M. Hoppe, M. Schlummer, and P. de Voogt. 2013. Presence and sources of anthropogenic perfluoroalkyl acids in high-consumption tap-water based beverages. Chemosphere 90:36-41. http://dx.doi.Org/10.1016/i.chemosphere.2012.06.070. Falandysz, J., S. Taniyasu, A. Gulkowska, N. Yamashita, and U. Schulte-Oehlmann. 2006. Is fish a major source of fluorinated surfactants and repellents in humans living on the Baltic Coast? Environmental Science and Technology 40:748-751. http://dx.doi.org/10.1021/esQ51799n. 83 ------- Fraser, A.J., T.F. Webster, D.J. Watkins, M.J. Strynar, K. Kato, A.M. Calafat, V.M. Vieira, and M.D. McClean. 2013. Polyfluorinated compounds in dust from homes, offices, and vehicles as predictors of concentrations in office workers' serum. Environment International 60:128-136. https://pubmed.ncbi.nlm.nih.gov/24041736/. Gebbink, W.A., A. Glynn, P.O. Darnerud, and U. Berger. 2015. Perfluoroalkyl acids and their precursors in Swedish food: The relative importance of direct and indirect dietary exposure. Environmental Pollution 198:108-115. http://dx.doi.Org/10.1016/i.envpol.2014.12.022. Giovanoulis, G., M.A. Nguyen, M. Arwidsson, S. Langer, R. Vestergren, and A. Lagerqvist. 2019. Reduction of hazardous chemicals in Swedish preschool dust through article substitution actions. Environment International 130:104921. http://dx.doi.Org/10.1016/i.envint.2019.104921. Groffen, T., M. Eens, and L. Bervoets. 2019. Do concentrations of perfluoroalkylated acids (PFAAs) in isopods reflect concentrations in soil and songbirds? A study using a distance gradient from a fluorochemical plant. Science of the Total Environment 657:111-123. https://pubmed.ncbi.nlm.nih.gov/30537574/. Gr0nnestad, R., B.P. Vazquez, A. Arukwe, V.L.B. Jaspers, B.M. Jenssen, M. Karimi, J.L. Lyche, and A. Kr0kje. 2019. Levels, patterns, and biomagnification potential of perfluoroalkyl substances in a terrestrial food chain in a Nordic skiing area. Environmental Science & Technology 53:13390-13397. https://pubmed.ncbi.nlm.nih.gov/31691564/. Hale, S.E., H.P. Arp, G.A. Slinde, E.J. Wade, K. Bj0rseth, G.D. Breedveld, B.F. Straith, K.G. Moe, M. Jartun, and A. H0isaeter. 2017. Sorbent amendment as a remediation strategy to reduce PFAS mobility and leaching in a contaminated sandy soil from a Norwegian firefighting training facility. Chemosphere 171:9-18. https://www.sciencedirect.eom/science/a rticle/abs/pii/S0045653516317775. Harrad, S., N. Wemken, D.S. Drage, M.A.E. Abdallah, and A.M. Coggins. 2019. Perfluoroalkyl substances in drinking water, indoor air and dust from Ireland: Implications for human exposure. Environmental Science & Technology 53:13449-13457. http://dx.doi.org/10.1021/acs.est.9b04604. Harrad, S., D.S. Drage, M. Sharkey, and H. Berresheim. 2020. Perfluoroalkyl substances and brominated flame retardants in landfill-related air, soil, and groundwater from Ireland. Science of the Total Environment 705:135834. http://dx.doi.Org/10.1016/i.scitotenv.2019.135834. Haug, L.S., S. Huber, M. Schlabach, G. Becher, and C. Thomsen. 2011. Investigation on per- and polyfluorinated compounds in paired samples of house dust and indoor air from Norwegian homes. Environmental Science & Technology 45:7991-7998. http://dx.doi.org/10.1021/eslQ3456h. 84 ------- Herzke, D., S. Huber, L. Bervoets, W. D'Hollander, J. Hajslova, J. Pulkrabova, G. Brambilla, S.P. De Filippis, S. Klenow, G. Heinemeyer, and P. de Voogt. 2013. Perfluorinated alkylated substances in vegetables collected in four European countries; occurrence and human exposure estimations. Environmental Science and Pollution Research 20:7930-7939. http://dx.doi.org/10.1007/sll356-013-1777-8. Hlouskova, V., P. Hradkova, J. Poustka, G. Brambilla, S.P. De Filipps, W. D'Hollander, L. Bervoets, D. Herzke, S. Huber, P. de Voogt, and J. Pulkrabova. 2013. Occurrence of perfluoroalkyl substances (PFASs) in various food items of animal origin collected in four European countries. Food Additives and Contaminants: Part A 30:1918-1932. http://dx.doi.org/10.1080/19440049.2Q13.837585. Holzer, J., T. Goen, P. Just, R. Reupert, K. Rauchfuss, M. Kraft, J. Miiller, and M. Wilhelm. 2011. Perfluorinated compounds in fish and blood of anglers at Lake Mohne, Sauerland area, Germany. Environmental Science and Technology 45:8046-8052. http://dx.doi.org/10.1021/eslQ4391z. Huber, S., L.S. Haug, and M. Schlabach. 2011. Per- and polyfluorinated compounds in house dust and indoor air from northern Norway—a pilot study. Chemosphere 84:1686-1693. http://dx.doi.Org/10.1016/i.chemosphere.2011.04.075. Jogsten, I.E., G. Perello, X. Llebaria, E. Bigas, R. Martf-Cid, A. Karrman, and J.L. Domingo. 2009. Exposure to perfluorinated compounds in Catalonia, Spain, through consumption of various raw and cooked foodstuffs, including packaged food. Food and Chemical Toxicology 47:1577-1583. http://dx.doi.Org/10.1016/i.fct.2009.04.004. Jogsten, I.E., M. Nadal, B. van Bavel, G. Lindstrom, and J.L. Domingo. 2012. Per- and polyfluorinated compounds (PFCs) in house dust and indoor air in Catalonia, Spain: Implications for human exposure. Environment International 39:172-180. http://dx.doi.Org/10.1016/i.envint.2011.09.004. Jorundsdottir, H., T.I. Halldorsson, and H. Gunnlaugsdottir. 2014. PFAAs in fish and other seafood products from Icelandic waters. Journal of Environmental and Public Health 2014:573607. http://dx.doi.org/10.1155/2014/573607. Karaskova, P., M. Venier, L. Melymuk, J. Becanova, S. Vojta, R. Prokes, M.L. Diamond, and J. Klanova. 2016. Perfluorinated alkyl substances (PFASs) in household dust in Central Europe and North America. Environment International 94:315-324. https://doi.Org/10.1016/i.envint.2016.05.031. Kato, K., A.M. Calafat, and L.L. Needham. 2009. Polyfluoroalkyl chemicals in house dust. Environmental Research 109:518-523. http://dx.doi.Org/10.1016/i.envres.2009.01.005. 85 ------- Knobeloch, L., P. Imm, and H. Anderson. 2012. Perfluoroalkyl chemicals in vacuum cleaner dust from 39 Wisconsin homes. Chemosphere 88:779-783. https://doi.Org/10.1016/i.chemosphere.2012.03.082. Kubwabo, C., B. Stewart, J. Zhu, and L. Marro. 2005. Occurrence of perfluorosulfonates and other perfluorochemicals in dust from selected homes in the city of Ottawa, Canada. Journal of Environmental Monitoring 7:1074-1078. http://dx.doi.org/10.1039/b5Q7731c. Lankova, D., 0. Lacina, J. Pulkrabova, and J. Hajslova. 2013. The determination of perfluoroalkyl substances, brominated flame retardants and their metabolites in human breast milk and infant formula. Talanta 117:318-325. http://dx.doi.Org/10.1016/i.talanta.2013.08.040. Mejia-Avendano, S., G. Munoz, S. Vo Duy, M. Desrosiers, P. Benoit, S. Sauve, and J. Liu. 2017. Novel fluoroalkylated surfactants in soils following firefighting foam deployment during the Lac-Megantic railway accident. Environmental Science & Technology 51:8313-8323. http://dx.doi.org/10.1021/acs.est.7b02028. Nickerson, A., A.C. Maizel, P.R. Kulkarni, D.T. Adamson, J.J. Kornuc, and C.P. Higgins. 2020. Enhanced extraction of AFFF-associated PFASs from source zone soils. Environmental Science & Technology 54:4952-4962. http://dx.doi.org/10.1021/acs.est.0cQ0792. Noorlander, C.W., S.P.J, van Leeuwen, J.D. te Biesebeek, M.J.B. Mengelers, and M.J. Zeilmaker. 2011. Levels of perfluorinated compounds in food and dietary intake of PFOS and PFOA in the Netherlands. Journal of Agricultural and Food Chemistry 59:7496-7505. http://dx.doi.org/10.1021/iflQ4943p. Padilla-Sanchez, J.A., and L.S. Haug. 2016. A fast and sensitive method for the simultaneous analysis of a wide range of per- and polyfluoroalkyl substances in indoor dust using on- line solid phase extraction-ultrahigh performance liquid chromatography-time-of-flight- mass spectrometry. Journal of Chromatography A 1445:36-45. http://dx.doi.Org/10.1016/i.chroma.2016.03.058. Papadopoulou, E., S. Poothong, J. Koekkoek, L. Lucattini, J.A. Padilla-Sanchez, M. Haugen, D. Herzke, S. Valdersnes, A. Maage, I.T. Cousins, P.E.G. Leonards, and L.S. Haug. 2017. Estimating human exposure to perfluoroalkyl acids via solid food and drinks: Implementation and comparison of different dietary assessment methods. Environmental Research 158:269-276. http://dx.doi.Org/10.1016/i.envres.2017.06.011. Perez, F., M. Llorca, M. Kock-Schulmeyer, B. Skrbic, L.S. Oliveira, K. da Boit Martinello, N.A. Al- Dhabi, I. Antic, M. Farre, and D. Barcelo. 2014. Assessment of perfluoroalkyl substances in food items at global scale. Environmental Research 135:181-189. http://dx.doi.Org/10.1016/i.envres.2014.08.004. 86 ------- Riviere, G., J. Jean, S. Gorecki, M. Hulin, M. Kolf-Clauw, C. Feidt, N. Picard-Hagen, P. Vasseur, B. Le Bizec, and V. Sirot. 2019. Dietary exposure to perfluoroalkyl acids, brominated flame retardants and health risk assessment in the French infant total diet study. Food and Chemical Toxicology 131:110561. http://dx.doi.Org/10.1016/i.fct.2019.06.008. Schecter, A., J. Colacino, D. Haffner, K. Patel, M. Opel, O. Papke, and L. Birnbaum. 2010. Perfluorinated compounds, polychlorinated biphenyls, and organochlorine pesticide contamination in composite food samples from Dallas, Texas, USA. Environmental Health Perspectives 118:796-802. http://dx.doi.org/10.1289/ehp.09Q1347. Scher, D.P., J.E. Kelly, C.A. Huset, K.M. Barry, R.W. Hoffbeck, V.L. Yingling, and R.B. Messing. 2018. Occurrence of perfluoroalkyl substances (PFAS) in garden produce at homes with a history of PFAS-contaminated drinking water. Chemosphere 196:548-555. http://dx.doi.Org/10.1016/i.chemosphere.2017.12.179. Scher, D.P., J.E. Kelly, C.A. Huset, K.M. Barry, and V.L. Yingling. 2019. Does soil track-in contribute to house dust concentrations of perfluoroalkyl acids (PFAAs) in areas affected by soil or water contamination? Journal of Exposure Science and Environmental Epidemiology 29:218-226. http://dx.doi.org/10.1038/s41370-018-Q101-6. Scordo, C.V.A., L. Checchini, L. Renai, S. Orlandini, M.C. Bruzzoniti, D. Fibbi, L. Mandi, N. Ouazzani, and M. Del Bubba. 2020. Optimization and validation of a method based on QuEChERS extraction and liquid chromatographic-tandem mass spectrometric analysis for the determination of perfluoroalkyl acids in strawberry and olive fruits, as model crops with different matrix characteristics. Journal of Chromatography A 1621:461038. http://dx.doi.Org/10.1016/i.chroma.2020.461038. Skaar, J.S., E.M. Raeder, J.L. Lyche, L. Ahrens, and R. Kallenborn. 2019. Elucidation of contamination sources for poly- and perfluoroalkyl substances (PFASs) on Svalbard (Norwegian Arctic). Environmental Science and Pollution Research International 26:7356-7363. https://pubmed.ncbi.nlm.nih.gov/29754295/. Strynar, M.J., and A.B. Lindstrom. 2008. Perfluorinated compounds in house dust from Ohio and North Carolina, USA. Environmental Science & Technology 42:3751-3756. http://dx.doi.org/10.1021/es7032058. Surma, M., M. Piskula, W. Wiczkowski, and H. Zieliriski. 2017. The perfluoroalkyl carboxylic acids (PFCAs) and perfluoroalkane sulfonates (PFSAs) contamination level in spices. European Food Research and Technology 243:297-307. http://dx.doi.org/10.1007/sQ0217-016-2744-7. Sznajder-Katarzyriska, K., M. Surma, E. Cieslik, and W. Wiczkowski. 2018. The perfluoroalkyl substances (PFASs) contamination of fruits and vegetables. Food Additives & Contaminants: Part A, Chemistry, Analysis, Control, Exposure & Risk Assessment 35:1776-1786. http://dx.doi.org/10.1080/19440049.2018.15Q2477. 87 ------- Sznajder-Katarzyriska, K., M. Surma, W. Wiczkowski, and E. Cieslik. 2019. The perfluoroalkyl substance (PFAS) contamination level in milk and milk products in Poland. International Dairy Journal 96:73-84. http://dx.doi.Org/10.1016/i.idairyi.2019.04.008. Vassiliadou, I., D. Costopoulou, N. Kalogeropoulos, S. Karavoltsos, A. Sakellari, E. Zafeiraki, M. Dassenakis, and L. Leondiadis. 2015. Levels of perfluorinated compounds in raw and cooked Mediterranean finfish and shellfish. Chemosphere 127:117-126. http://dx.doi.Org/10.1016/i.chemosphere.2014.12.081. Venkatesan, A.K., and R.U. Halden. 2014. Loss and in situ production of perfluoroalkyl chemicals in outdoor biosolids-soil mesocosms. Environmental Research 132:321-327. http://dx.doi.Org/10.1016/i.envres.2014.04.024. Winkens, K., G. Giovanoulis, J. Koponen, R. Vestergren, U. Berger, A.M. Karvonen, J. Pekkanen, H. Kiviranta, and I.T. Cousins. 2018. Perfluoroalkyl acids and their precursors in floor dust of children's bedrooms—Implications for indoor exposure. Environment International 119:493-502. http://dx.doi.Org/10.1016/i.envint.2018.06.009. Young, W.M., P. South, T.H. Begley, G.W. Diachenko, and G.O. Noonan. 2012. Determination of perfluorochemicals in cow's milk using liquid chromatography-tandem mass spectrometry. Journal of Agricultural and Food Chemistry 60:1652-1658. http://dx.doi.org/10.1021/if2Q4565x. Young, W.M., P. South, T.H. Begley, and G.O. Noonan. 2013. Determination of perfluorochemicals in fish and shellfish using liquid chromatography—tandem mass spectrometry. Journal of Agricultural and Food Chemistry 61:11166-11172. http://dx.doi.org/10.1021/if4Q3935g. Zafeiraki, E., D. Costopoulou, I. Vassiliadou, L. Leondiadis, E. Dassenakis, R.L.A.P. Hoogenboom, and S.P.J, van Leeuwen. 2016a. Perfluoroalkylated substances (PFASs) in home and commercially produced chicken eggs from the Netherlands and Greece. Chemosphere 144:2106-2112. http://dx.doi.Org/10.1016/i.chemosphere.2015.10.105. Zafeiraki, E., I. Vassiliadou, D. Costopoulou, L. Leondiadis, H.A. Schafft, R.L.A.P. Hoogenboom, and S.P.J, van Leeuwen. 2016b. Perfluoroalkylated substances in edible livers of farm animals, including depuration behaviour in young sheep fed with contaminated grass. Chemosphere 156:280-285. http://dx.doi.Org/10.1016/i.chemosphere.2016.05.003. Zheng, G., B.E. Boor, E. Schreder, and A. Salamova. 2020. Indoor exposure to per- and polyfluoroalkyl substances (PFAS) in the childcare environment. Environmental Pollution 258:113714. http://dx.doi.Org/10.1016/i.envpol.2019.113714. 88 ------- Appendix D: Comparative Analysis for Potentially Sensitive Populations for PFBS The EPA evaluated several exposure scenarios for PFBS to determine whether the national recommended criteria for the general population, male and female adults > 21 years old, are sufficiently protective of potentially sensitive subpopulations. To accomplish this, the EPA considered three additional exposure scenarios, as supported by data from the EPA Exposure Factors Handbook (EFH; EPA, 2011) and the Human Health Methodology (EPA, 2000). Specifically, the EPA evaluated exposure parameters for "all ages" as well as two potentially sensitive life stages associated with the critical effect used to derive the PFBS chronic RfD, i.e., adverse developmental effect on thyroid activity, specifically decreased serum total thyroxine, in newborn mice (postnatal day [PND] 1) born to mothers that had been orally exposed to K+PFBS throughout gestation (EPA, 2021a,b). Based on this exposure interval in the critical study, potentially sensitive subpopulations in humans include women of childbearing age who may be or become pregnant and pregnant women (Table D-l). For the body weight exposure parameter, a mean bodyweight of 75 kg for pregnant women (all trimesters) was identified in the EFH (2011, Ch. 8, Table 8-29). A representative body weights for the "all ages" scenario was not specifically presented in the EFH (EPA, 2011). To address this data limitation, for this exercise, the EPA assumed that the average body weight for "all ages" was 71.6 kg based on the sum of the time-weighted averages of the mean male and female combined body weights from 1 year up to 80 years old from the NHANES (1999-2006) (Table 8- 3; EPA, 2011). A body weight average of 67 kg for women of childbearing age was identified in the Human Health Methodology (EPA, 2000); however, this average is based on an older NHANES dataset (NHANES III; WESTAT 2000). More recent NHANES data (1999-2006) suggest that the mean body weight for women of childbearing age ranges from 65.9 kg for 16 to < 21- year-olds to 77.1 kg for 40 to < 50-year-olds (Table 8-5; EPA, 2011). Using these data, the EPA assumed a time-weighted average body weight of 73.4 kg for women of childbearing age (Table 8-5; EPA, 2011). Drinking water intake values were available for all populations (Table D-l). The EPA encountered several data limitations for trophic level specific fish consumption rates for some of these potentially sensitive populations. The EPA's national criteria are typically derived using trophic-level specific fish consumption rates (FCRs), paired with trophic-level specific bioaccumulation factors (BAFs) to account for the potential bioaccumulation of some chemicals in aquatic food webs and the broad physiological differences between trophic levels which may influence bioaccumulation (EPA, 2000). Trophic level specific FCRs for women of childbearing age were identified (Table D-l). However, trophic level specific FCRs are not available for two of the potentially sensitive life stages: all ages and pregnant women. Therefore, criteria could not be calculated for these life stages. However, in all cases with available data, the total FCR for the alternative scenarios is lower than the FCR for the general population. Because bodyweights are similar for all of the considered populations (see above and Table D-l), the FCR is likely to be the main determinant of the criteria value, with a larger FCR resulting in a lower, more health protective criterion. Therefore, criteria based on the general population are expected to be protective of the identified potentially sensitive life 89 ------- Table D-l. Comparison of noncancer-based HHC values for different candidate sensitive populations identified from the critical effect and study. Population Bodyweight (kg) Drinking Water Intake (L/day) Fish Consun (g/d iption Rate ay) Criteria (Hg/L) Total TL 2 TL 3 TL 4 W + O 00 General, adult (> 21 years) 80a 2.3b 22° 7.6° 8.6° 5.1c 0.4 0.5 Women of childbearing Age (13-49 years) 73.4d 2.1e 15.8C 5.6C 6.0C 2.9C 0.6 0.8 All Ages (Birth to 80 years) 71.6f 2.0b 19.3g NA NA NA ND ND Pregnant Women 75h 2.1e 10' NA NA NA ND ND Notes: g/day = grams offish consumed per day; L/day = liters of water per day; NA = not available; ND = not determined; 00 = organism only; W + O = water plus organism. Bold values indicate draft national recommended criteria. Gray highlighting indicates most health protective HHC based on noncancer effects. a EPA, 2011, Exposure Factors Handbook, Ch. 8, Table 8-1, NHANES 1999-2006.Recommended mean bodyweight for adults. b Estimated using the FCID calculator (University of Maryland, 2024; https://fcid.foodrisk.org/), NHANES 2005- 2010, community water, 90th percentile per capita rate. c EPA, 2014; NHANES 2003-2010 survey data, 90th percentile per capita rate, freshwater and estuarine fish and shellfish edible portion, adults > 21 years. dTime weighted average of combined bodyweights for women ages 16 to < 50 years, NHANES 1999-2006 (EPA, 2011; Table 8-5). e EPA, 2019, Exposure Factors Handbook; Update Ch. 3., Table 3-62, Community water, 90th percentile, per capita rate. fTime weighted average of mean male and female combined body weights from 1 year up to 80 years, NHANES 1999-2006 (EPA, 2011; Table 8-3). g Estimated using the FCID calculator (University of Maryland, 2024; https://fcid.foodrisk.org/), NHANES 2005- 2010; freshwater and estuarine fish and shellfish combined, 90th percentile per capita rate; male and female, all ages included. h EPA, 2011, Exposures Factors Handbook, Ch 8, mean, NHANES 1999-2006, Table 8-29 i Estimated using the FCID calculator (University of Maryland, 2024; https://fcid.foodrisk.org/), NHANES 2005- 2010; freshwater and estuarine fish and shellfish combined, 90th percentile per capita rate pregnant females only. stages (Table D-l). Separately, paired bodyweight adjusted FCRs are not available for specific trophic levels which precludes the use of body-weight adjusted DWI rates to derive ambient water quality criteria. For illustrative purposes, the EPA calculated criteria based on the exposure parameters for women of childbearing age. As demonstrated in Table D-l, criteria based on the exposure inputs for the general population result in more health protective criteria and thus are protective of the potentially susceptible life stage of women of childbearing age (Table D-l). Overall, when bodyweight averages are similar, the resulting criteria are driven predominantly by the FCR; thus, a higher FCR results in a more health protective criteria. 90 ------- References EPA (Environmental Protection Agency). 2000. Methodology for Deriving Ambient Water Quality Criteria for the Protection of Human Health (2000). EPA-822-B-00-004. EPA, Office of Water, Office of Science and Technology, Washington, DC. Accessed January 2024. https://www.epa.gov/sites/default/files/2018-10/documents/methodology-wqc- protection-hh-2000.pdf. EPA (Environmental Protection Agency). 2011. Body Weight Studies. Chapter 8 in Exposure Factors Handbook. EPA/600/R-09/052F. EPA, National Center for Environmental Assessment, Office of Research and Development, Washington, DC. Accessed August 2024. https://www.epa.gov/sites/default/files/2015-09/documents/efh-chapter08.pdf. EPA (Environmental Protection Agency). 2014. Estimated Fish Consumption Rates for the U.S. Population and Selected Subpopulations (NHANES 2003-2010). EPA-820-R-14-002. EPA. Accessed August 2024. https://www.epa.gov/sites/default/files/2015-01/documents/fish-consumption-rates- 2014.pdf. EPA (Environmental Protection Agency). 2019. Update for Chapter 3 of the Exposure Factors Handbook, Ingestion of Water and Other Select Liquids. EPA/600/R-18/259F. EPA, Office of Research and Development, National Center for Environmental Assessment, Washington, DC. Accessed August 2024. https://www.epa.gov/sites/default/files/2019-02/documents/efh - chapter 3 update.pdf. EPA (Environmental Protection Agency). 2021a. Provisional Peer-Reviewed Toxicity Values for Perfluorobutane Sulfonic Acid (PFBS) and Related Compound Potassium Perfluorobutane Sulfonate. EPA/690/R-21/001. EPA, Office of Research and Development, Center for Public Health and Environmental Assessment, Cincinnati, OH. Accessed February 2024. https://cfpub.epa.gov/ncea/pprtv/recordisplay.cfm?deid=350061. EPA (Environmental Protection Agency). 2021b. Human Health Toxicity Values for Perfluorobutane Sulfonic Acid (CASRN 375-73-5) and Related Compound Potassium Perfluorobutane Sulfonate (CASRN 29420-49-3). EPA-600-R-20-345F. EPA, Office of Research and Development, Washington, DC. Accessed February 2024. https://www.epa.gov/pfas/learn-about-human-health-toxicity-assessment-pfbs. University of Maryland. 2024. What We Eat in America—Food Commodity Intake Database 2005-10. University of Maryland, College Park, MD, and EPA Office of Pesticide Programs, Washington, DC. Accessed August 2024. https://fcid.foodrisk.org/. WESTAT. 2000. Memorandum on Body Weight Estimates Based on NHANES III Data, Including Data Tables and Graphs. Analysis Conducted and Prepared by WESTAT under EPA Contract No. 68-C-99-242. March 3, 2000. 91 ------- |