|-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


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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.


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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

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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

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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.


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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

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(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

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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.

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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.

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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).

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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.

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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.

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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.

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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).

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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,

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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.

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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.

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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.

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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.

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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).

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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)

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•	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.

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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).

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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,

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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;

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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

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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.

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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

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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.

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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.

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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

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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

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(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).

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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

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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).

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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)

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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)

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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

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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

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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

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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

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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

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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

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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

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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


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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


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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.

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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)

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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.

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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


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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


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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


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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


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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


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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


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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


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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%.

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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


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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


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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


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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


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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

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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.

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References

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Criteria for the Protection of Human Health (2000). EPA-822-B-00-004. EPA, Office of
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EPA (Environmental Protection Agency). 2011. Body Weight Studies. Chapter 8 in Exposure
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