&EPA

United States	Office of Water

Environmental Protection OST
Agency	EPA

PUBLICATION #
EPA/822/R-22/006

Drinking Water Health Advisory:
Perfluorobutane Sulfonic Acid (CASRN 375-73-5) and
Related Compound Potassium Perfluorobutane Sulfonate

(CASRN 29420-49-3)


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Drinking Water Health Advisory:

Perfluorobutane Sulfonic Acid (CASRN 375-73-5) and Related Compound
Potassium Perfluorobutane Sulfonate (CASRN 29420-49-3)

Prepared by:

U.S. Environmental Protection Agency
Office of Water (4304T)

Office of Science and Technology
Health and Ecological Criteria Division
Washington, DC 20460

EPA Document Number: EPA/822/R-22/006
June 2022


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Acknowledgments

This document was prepared by the Health and Ecological Criteria Division (HECD), Office of
Science and Technology (OST), Office of Water (OW) of U.S. Environmental Protection
Agency (EPA).

OW scientists and managers who provided valuable contributions and direction in the
development of this health advisory are, from OST: Czarina Cooper, MPH (lead); Casey
Lindberg, PhD; Carlye Austin, PhD, DABT; Kelly Cunningham, MS; Brittany Jacobs, PhD;
Susan Euling, PhD; and Colleen Flaherty, MS; and from the Office of Ground Water and
Drinking Water (OGWDW): Stanley Gorzelnik, PE; Daniel P. Hautman; Ashley Greene, MS;
and Ryan Albert, PhD.

The agency gratefully acknowledges the valuable contributions of EPA scientists Jason C.
Lambert, PhD, DABT, and Elizabeth Oesterling Owens, PhD, from the Office of Research and
Development (ORD).

The literature searches to identify information about the relative source contribution for PFBS
were performed by contractors at ICF (contract number 68HE0C18D0001) and Tetra Tech
(contract number 68HERC20D0016).

This document was provided for review by staff in the following EPA Program Offices and
Regions:

•	Office of Water

•	Office of Chemical Safety and Pollution Prevention, Office of Pollution Prevention and
Toxics

•	Office of Land and Emergency Management

•	Office of Policy

•	Office of Children's Health Protection

•	Office of Research and Development

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Contents

Abbreviations and Acronyms	iv

Executive Summary	vii

1.0 Introduction and Background	1

1.1	History under SDWA	1

1.2	Current Advisories and Guidelines	2

1.3	Uses and Sources of PFBS	4

1.4	Environmental Fate, Occurrence in Water, and Exposure to Humans	4

1.4.1	Environmental Fate and Transport in the Environment	4

1.4.2	Occurrence in Water	5

1.4.3	Exposure in Humans	11

2.0 Problem Formulation and Scope	11

2.1	Conceptual Model	11

2.2	Analysis Plan	13

2.2.1	Health Advisory Guidelines	13

2.2.2	Sources of Toxicity Information for Health Advisory Development	14

2.2.3	Approach and Scope for Health Advisory Derivation	15

2.2.4	Exposure Factors for Deriving Health Advisory	17

3.0 Health Advisory Input Values	20

3.1	Toxicity Assessment Values	20

3.2	Exposure Factors	21

3.3	Relative Source Contribution	22

3.3.1	Non-Drinking Water Sources and Routes	22

3.3.2	RSC Determination	31

4.0 Lifetime Noncancer Health Advisory Derivation	32

5.0 Analytical Methods	32

6.0 Treatment Technologies	33

6.1	Point-of-Use Devices for Individual Household PFBS Removal	35

6.2	Treatment Technologies Summary	35

7.0 Consideration of Noncancer Health Risks from PFAS Mixtures	36

8.0 Health Advisory Characterization	37

8.1 Comparative Analysis of Exposure Factors for Different Populations	38

9.0 References	39

Appendix A: Relative Source Contribution - Literature Search and Screening

Methodology	67

Appendix B: Compilation of Data on PFBS Occurrence in Environmental Media

Collected from Primary Literature	70

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Figures

Figure 1. Conceptual Model for PFBS Drinking Water Health Advisory Development	12

Tables

Table 1. State Guideline Values for PFBS	2

Table 2. International Guideline Values for PFBS	3

Table 3. EPA Exposure Factors for Drinking Water Intake	18

Table 4. Chronic Noncancer Toxicity Information for PFBS for Deriving the Lifetime HA	20

Table 5. EPA Exposure Factors for Drinking Water Intake for Different Candidate

Sensitive Populations or Life Stages Based on the Critical Effect and Study	22

Table 6. Comparison of HA Values Using EPA Exposure Factors for Drinking Water

Intake for Different Candidate Populations	38

Table A-l. Populations, Exposures, Comparators, and Outcomes (PECO) Criteria	68

Table B-l. Compilation of Studies Describing PFBS Occurrence in Drinking Water	70

Table B-2. Compilation of Studies Describing PFBS Occurrence in Groundwater	77

Table B-3. Compilation of Studies Describing PFBS Occurrence in Surface Water	79

Table B-4. Compilation of Studies Describing PFBS Occurrence in Food	86

Table B-5. Compilation of Studies Describing PFBS Occurrence in Indoor Dust	91

Table B-6. Compilation of Studies Describing PFBS Occurrence in Soil	93

in


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Abbreviations and Acronyms

6:2 FTS	6:2 fluorotelomer sulfonic

acid

ADAF

age-dependent adjustment



factor

AFFF

aqueous film-forming foam

ANSI

American National



Standards Institute

ASTSWMO

Association of State and



Territorial Solid Waste



Management Officials

AT SDR

Agency for Toxic



Substances and Disease



Registry

BMD

benchmark dose

BMDL

benchmark dose lower limit

bw or BW

body weight

CDPHE

Colorado Department of



Public Health and



Environment

CPHEA

Center for Public Health



and Environmental



Assessment

CSF

cancer slope factor

DF

detection frequency

DOH

Department of Health

DQO

data quality objective

dw

dry weight

DWI

drinking water intake

DWI-BW

body weight-adjusted



drinking water intake

DWTP

drinking water treatment



plant

ECHA

European Chemicals



Agency

EEE

electronic equipment

EF

exposure factor

EFH

Exposure Factors



Handbook

EFSA

European Food Safety



Authority

EGLE

Michigan Department of



Environment, Great Lakes,



and Energy

EPA

United States



Environmental Protection



Agency

Eq

equation

EU

European Union

FCID

Food Commodity Intake



Database

FCM

food contact material

FDA

United States Food and



Drug Administration

fw

fresh weight

GCA

groundwater contamination



area

GenX chemicals

; hexafluoropropylene oxide



dimer acid and its



ammonium salt

HA

Health Advisory

HED

human equivalent dose

HI

hazard index

HIDOH

Hawai i Department of



Health

HQ

hazard quotient

IBWA

International Bottled Water



Association

IDEM

Indiana Department of



Environmental



Management

iHA

Interim Health Advisory

Illinois EPA

Illinois Environmental



Protection Agency

Kaw

air-water partition



coefficient

km2

square kilometers

K+PFBS

potassium perfluorobutane



sulfonate

L

liters

L/kg bw-day

liters per kilogram body



weight per day

IV


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L/m2hr

liters per square meter per

NRSA

National Rivers and



hour



Streams Assessment

LC/MS/MS

liquid







chromatography/tandem

ODH

Ohio Department of Health



mass spectrometry

OEHHA

Office of Environmental

LOAEL

lowest observed adverse



Health Hazard Assessment



effect level

Ohio EPA

Ohio Environmental

LOQ

limit of quantification



Protection Agency

Maine DEP

Maine Department of

ORD

Office of Research and



Environmental Protection



Development

MCLG

Maximum Contaminant

OST

Office of Science and



Level Goal



Technology

MDH

Minnesota Department of

OW

Office of Water



Health

PECO

populations, exposures,

mg/kg bw-day

milligrams per kilogram



comparators, and outcomes



body weight per day

PFAA

perfluoroalkylated acid

MPa

megapascal

PFAS

per- and polyfluoroalkyl

MRL

minimum reporting level



substances

MS/MS

tandem mass spectrometry

PFBA

perfluorobutanoic acid

MW

molecular weight

PFBS

perfluorobutane sulfonic

NCHS

National Center for Health



acid



Statistics

PFC

perfluorinated chemical

NCOD

National Contaminant

PFCA

perfluoroalkyl carboxylic



Occurrence Database



acid

ND

non-detect

PFDA

perfluorodecanoic acid

NDEP

Nevada Division of

PFDoDA

perfluorododecanoic acid



Environmental Protection

PFDoS

perfluorododecane sulfonic

NF

nanofiltration



acid

ng/g

nanograms per gram

PFDS

perfluorodecane sulfonic

ng/kg bw-day

nanograms per kilogram



acid



body weight per day

PFHpA

perfluoroheptanoic acid

ng/L

nanograms per liter

PFHpS

perfluoroheptane sulfonic

ng/mL

nanograms per milliliter



acid

NHANES

National Health and

PFHxA

perfluorohexanoic acid



Nutrition Examination

PFHxS

perfluorohexane sulfonic



Survey



acid

NHIS

Human Nutrition

PFNA

perfluorononanoic acid



Information Service

PFNS

perfluorononane sulfonic

NO A A

National Oceanic and



acid



Atmospheric

PFOA

perfluorooctanoic acid



Administration

PFOS

perfluorooctanesulfonic

NOAEL

no observed adverse effect



acid



level

PFOSA

perfluorooctanesulfonami de

V


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PFPA

perfluoropentanoic acid

PFPeA

perfluoropentanoic acid

PFPS

perfluoropentane sulfonic



acid

PFSA

perfluoroalkane sulfonate

PFTrDA

perfluorotridecanoic acid

PFTrS

perfluorotridecane sulfonic



acid

PFUnDA

perfluoroundecanoic acid

PFUnS

perfluoroundecane



sulfonate

pg/cm2

picograms per square



centimeter

pg/m3

picograms per cubic meter

PM

particulate matter

PMN

pre-manufacture notice

PND

post-natal day

POD

point of departure

ppm

parts per million

ppt

parts per trillion

POE

point-of-entry

POU

point-of-use

PWS

public water system

QC

quality control

REACH

Registration, Evaluation,



Authorization and



Restriction of Chemicals

RfD

reference dose

RO

reverse osmosis

RPF

relative potency factor

RSC

relative source contribution

SDWA

Safe Drinking Water Act

t4

thyroxine

TCEQ

Texas Commission on



Environmental Quality

TSCA

Toxic Substances Control

Act

UCMR	Unregulated Contaminant

Monitoring Rule
UCMR 3 third Unregulated

Contaminant Monitoring
Rule

UCMR 5

fifth Unregulated



Contaminant Monitoring



Rule

UF

uncertainty factor(s)

UFa

interspecies uncertainty



factor

UFc

composite uncertainty



factor

UFd

database uncertainty factor

UFh

intraspecies uncertainty



factor

UFl

LOAEL to NOAEL



extrapolation uncertainty



factor

UFs

extrapolation from



sub chronic to chronic



exposure duration



uncertainty factor

WEEE

wastes of electrical and



electronic equipment

Wisconsin DHS

Wisconsin Department of



Health Services

WTP

water treatment plant

WW

wet weight

WWTP

wastewater treatment plant

Mg/kg

micrograms per kilogram

|ig/m2

micrograms per square



meter

VI


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Executive Summary

Perfluorobutane sulfonic acid (PFBS; CASRN 375-73-5) and its related compound potassium
perfluorobutane sulfonate (K+PFBS; CASRN 29420-49-3) are shorter-chain members of a group
of substances known as per- and polyfluoroalkyl substances (PFAS). In water, K+PFBS fully
dissociates to the deprotonated anionic form of PFBS (PFBS-; CASRN 45187-15-3) and the K+
cation at environmental pH levels (pH 4-9). Herein, these three PFBS chemical forms are
referred to collectively as PFBS.

PFBS is a replacement chemical for the longer-chain perfluorooctane sulfonic acid (PFOS), a
PFAS that was voluntarily phased out (with some exceptions) by its primary U.S. manufacturer
(3M Company) between 2000 and 2002 (U.S. EPA, 2007; 3M, 2002). Prior to its use as a
replacement for PFOS, PFBS was produced as a byproduct during production of perfluorooctane
sulfonyl fluoride-based chemicals and was present in consumer products as an impurity
(AECOM, 2019). PFBS is used in the manufacture of paints, cleaning agents, and water- and
stain-repellent products and coatings (U.S. EPA, 2021a). PFBS has been detected in drinking
water, groundwater, and surface water and has been found in dust, carpeting and carpet cleaners,
floor wax, foods including seafood (fish and shellfish) and vegetables, food packaging, indoor
and outdoor air, soil, biosolids, and some consumer products (ATSDR, 2021; U.S. EPA, 2021a;
see Section 3.3.1). PFBS can enter the aquatic environment through releases from manufacturing
sites, industrial uses, fire/crash training areas, and wastewater treatment facilities, as well as from
land application of contaminated biosolids (ATSDR, 2021; U.S. EPA, 2021a). PFBS is water
soluble (52.6 g/L at 22.5-24 °C for the potassium salt) and volatilization from water surfaces is
not expected to be an important fate process (ATSDR, 2021; U.S. EPA, 2021a). PFBS has been
detected in the serum of humans in the general population (U.S. EPA, 2021a).

The U.S. Environmental Protection Agency (EPA) is issuing a lifetime noncancer drinking water
Health Advisory (HA) for PFBS of 2,000 nanograms per liter (ng/L) or 2,000 parts per trillion
(ppt). This is the first HA for PFBS and its finalization fulfills a commitment described in EPA's
PFAS Strategic Roadmap (U.S. EPA, 2021b). The final PFBS toxicity assessment titled Human
Health Toxicity Values for Perfluorobutane Sulfonic Acid (CASRN 375-73-5) and Related
Compound Potassium Perfluorobutane Sulfonate (CASRN 29420-49-3) (U.S. EPA, 2021a)
serves as the basis of the toxicity information used to derive the lifetime noncancer HA for
PFBS. The critical adverse effect is thyroid effects in mice (specifically, decreased serum levels
of the thyroid hormone thyroxine [T4]) observed at post-natal day (PND) 1, after 20-day
gestational exposure to PFBS (Feng et al., 2017). Based on this critical effect, a chronic
reference dose (RfD) of 3 x 10 4 milligrams per kilogram body weight per day (mg/kg bw-day)
for PFBS was derived.

In accordance with EPA's Recommended Use of Body Weight4 as the Default Method in
Derivation of the Oral Reference Dose (U.S. EPA, 2011), serum PFBS half-lives were used to
scale a toxicologically equivalent dose of orally administered PFBS from animals to humans.
Following EP A's Benchmark Dose Technical Guidance (U.S. EPA, 2012b), benchmark dose
(BMD) modeling of thyroid effects in offspring after gestational exposure to PFBS resulted in a
benchmark dose lower confidence limit (BMDL) for 0.5 SD change from the control
(BMDLo.5sd) human equivalent dose (HED) of 0.095 mg/kg bw-day. This HED point of
departure (POD) based on decreased levels of T4 in newborn offspring was divided by a
composite uncertainty factor (UFc) of 300 to derive the chronic RfD.

Vll


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Sensitive populations or life stages within the general population indicated by the critical study
used to derive the chronic RfD for PFBS are the developing embryo and fetus. Therefore,
drinking water exposure to pregnant women as well as women of childbearing age, who may be
or become pregnant, were identified as two sensitive populations or life stages. EPA selected the
body weight-adjusted drinking water intake (DWI-BW) exposure factor (EF) of 0.0354 liters per
kilogram body weight per day (L/kg bw-day) for women of childbearing age because it is more
health protective than the DWI-BW for pregnant women. However, PFBS HA values, when
rounded to one significant figure, were the same when calculated using EFs for either women of
childbearing age, pregnant women, or the general population (all ages).

The physical/chemical properties and available exposure information for PFBS suggest multiple
potentially significant exposure sources (seafood, other foods, indoor air, and some consumer
products) other than drinking water ingestion. However, information is not available to
quantitatively characterize the relative exposure contributions from non-drinking water exposure
sources. Therefore, following the Exposure Decision Tree approach within EPA's 2000
Methodology for Deriving Ambient Water Quality Criteria for the Protection of Human Health
(U.S. EPA, 2000a), EPA recommends a relative source contribution (RSC) of 20 percent (0.20)
for use in PFBS HA derivation.

There is insufficient toxicity information available to derive a one-day HA for PFBS. Derivation
of a 10-day HA was considered because the subchronic and chronic RfDs are both based on a
20-day exposure study, which may be used to derive a 10-day HA. However, EPA did not derive
a 10-day HA because the critical health effect on which the chronic RfD used to calculate the
lifetime HA is based (i.e., decreased serum levels of T/dn newborn mice) resulted from PFBS
exposure during a developmental life stage (Feng et al., 2017). EPA's risk assessment guidelines
for developmental toxicity indicate that adverse effects can result from even brief exposure
during a critical period of development (U.S. EPA, 1991). The critical study observed persistent
health effects into adulthood, suggesting the potential for long-term health consequences of
gestational-only PFBS exposure and that gestation is at least one critical exposure window for
PFBS. Therefore, the lifetime HA for PFBS of 2000 ng/L and the chronic RfD from which it is
derived are considered applicable to short-term PFBS exposure (including during pregnancy) as
well as lifetime exposure via drinking water. This lifetime HA applies to PFBS (CASRN 375-73-
5), K+PFBS (CASRN 29420-49-3), and PFBS- (CASRN 45187-15-3).

No studies evaluating the carcinogenicity of PFBS in humans or animals were identified (U.S.
EPA, 2021a). In accordance with EPA's Guidelines for Carcinogen Risk Assessment (U.S. EPA,
2005b), EPA concluded that there is "Inadequate Information to Assess Carcinogenic Potential"
for PFBS by any route of exposure (U.S. EPA, 2021a). Therefore, a 10"6 cancer risk
concentration cannot be derived for PFBS at this time.

EPA developed two analytical methods to quantitatively assess drinking water for targeted PFAS
that include PFBS: EPA Method 533 (U.S. EPA, 2019b), which has a quantitation limit of 3.5
ng/L for PFBS, and EPA Method 537.1, Version 2.0 (U.S. EPA, 2020b), which has a
quantitation limit of 6.3 ng/L for PFBS. These analytical methods can both effectively and
accurately measure PFBS in drinking water at levels significantly lower than the lifetime HA of
2,000 ng/L. EPA finished drinking water sampling results have not identified PFBS levels that
approached the lifetime HA of 2,000 ng/L. However, treatment technologies, including reverse
osmosis (RO), nanofiltration (NF), and sorption-based processes such as activated carbon and


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ion exchange are available and have been shown to remove PFBS in drinking water; however,
sorption has less efficacy with PFBS than similar longer-chained PFAS.

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1.0	Introduction and Background

The Safe Drinking Water Act (SDWA) (42 U.S.C. § § 300f - 300j-27) authorizes the U.S.
Environmental Protection Agency (EPA) to develop drinking water Health Advisories (HAs).1
HAs are national non-enforceable, non-regulatory drinking water concentration levels of a
specific contaminant at or below which exposure for a specific duration is not anticipated to lead
to adverse human health effects.2 HAs are intended to provide information that tribal, state, and
local government officials and managers of public water systems (PWSs) can use to determine
whether actions are needed to address the presence of a contaminant in drinking water. HA
documents reflect the best available science and include HA values as well as information on
health effects, analytical methodologies for measuring contaminant levels, and treatment
technologies for removing contaminants from drinking water. EPA's lifetime HAs identify levels
to protect all Americans, including sensitive populations and life stages, from adverse health
effects resulting from exposure throughout their lives to contaminants in drinking water.

In April 2021, EPA published a final toxicity assessment for two per- and polyfluoroalkyl
substances (PFAS): perfluorobutane sulfonic acid (PFBS) and its related compound potassium
perfluorobutane sulfonate (K+PFBS) (U.S. EPA, 2021a). K+PFBS differs from PFBS by being
associated with a potassium ion. In water, K+PFBS fully dissociates to the deprotonated anionic
form of PFBS (PFBS-; CASRN 45187-15-3) and the K+ cation at environmental pH levels (pH
4-9). Herein, these three PFBS chemical forms are referred to collectively as PFBS. Completing
the toxicity assessment was an essential step to better understanding the potential human health
effects of exposure to PFBS. The chronic noncancer reference dose (RfD) calculated in the
toxicity assessment allows EPA to develop a final lifetime HA that will help communities make
informed decisions to better protect human health. The final PFBS HA satisfies a commitment
described in EPA's PFAS Strategic Roadmap (U.S. EPA, 2021b).

1.1	History under SDWA

PFBS is not currently regulated under SDWA. The 1996 amendments to SDWA require that
EPA issue a new list of unregulated contaminants (once every five years) to be monitored by
PWSs.3 Under the Unregulated Contaminant Monitoring Rule (UCMR), EPA samples drinking
water systems to collect data for contaminants that are known or suspected to be found in
drinking water and do not have health-based standards under SDWA. The first four UCMRs
required monitoring of all large public drinking water systems (>10,000 people), and a subset of
smaller systems serving <10,000 people. PFBS was one of six PFAS monitored in drinking
water under the third UCMR (UCMR 3) between 2013 and 2015 (U.S. EPA, 2012a). It is also
one of 29 PFAS that will be monitored under the fifth UCMR (UCMR 5) between 2023 and
2025 (U.S. EPA, 2021c). The collection of drinking water occurrence data supports EPA's future

1	SDWA § 1412(b)(1)(F) authorizes EPA to "publish health advisories (which are not regulations) or take other appropriate
actions for contaminants not subject to any national primary drinking water regulation." www.epa.gov/sites/default/files/2020-
05/documents/safe_drinkiiig_water_act-title_xiv of public	_health_service_act.pdf

2	This document is not a regulation and does not impose legally binding requirements on EPA, states, tribes, or the regulated
community. This document is not enforceable against any person and does not have the force and effect of law. No part of this
document, nor the document as a whole, constitutes final agency action that affects the rights and obligations of any person. EPA
may change any aspects of this document in the future.

3	SDWA § 1445(a)(l )(D)(2)(B) provides that "Not later than 3 years after the date of enactment of the Safe Drinking Water Act
Amendments of 1996 and every 5 years thereafter, the Administrator shall issue a list pursuant to subparagraph (A) of not more
than 30 unregulated contaminants to be monitored by public water systems and to be included in the national drinking water
occurrence data base maintained pursuant to subsection (g)."

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regulatory determinations and may support additional actions to protect public health (U.S. EPA,
2021c).

1.2 Current Advisories and Guidelines

Table 1 provides final drinking water guideline values for PFBS that have been developed by
states. The state values range from 100 to 667,000 parts per trillion (ppt) or nanograms per liter
(ng/L); this broad range of values may in part reflect differences in the type of value derived,
state guidance/methodology for deriving values, or data included in the evaluation (see
references for more details).

Table 1. State Guideline Values for PFBS

State "'b

PFBS Level
(ppt |ng/L|)

Standard/Guidance

Type of Medium

Reference

California

500

Notification level

Drinking water

California OEHHA
(2021)

Colorado

400,000

Translation level

Groundwater;
Surface water

CDPHE (2020b)

Hawai'i

600

Environmental action levels

Groundwater

HIDOH (2021)

Illinois

2,100

Health-based guidance level

Drinking water;
Groundwater

Illinois EPA
(2021a)

Indiana

>2,100

Action level

Drinking water

IDEM (2022)

Maine

400,000

Remedial action guideline

Groundwater

Maine DEP (2018)

Michigan

420

Maximum contaminant level

Drinking water;
Groundwater

EGLE (2020)

Minnesota

100

Health-based value

Drinking water;
Groundwater

MDH (2022)

Nevada

667,000

Basic comparison level

Drinking water

NDEP (2020)

Ohio

2,100

Action level

Drinking water

Ohio EPA and
ODH (2022)

Pennsylvania

10,000

Medium-specific
concentration

Groundwater;
Residential use

Environmental
Quality Board
(2021)

29,000

Medium-specific
concentration

Groundwater; Non-
residential use

Texas

34,000

Tier 1 protective
concentration level

Groundwater

TCEQ (2021)

Washington

345

State action level

Drinking water

Washington DOH
(2021)

Wisconsin

450,000

Recommended enforcement
standard

Groundwater

Wisconsin DHS
(2020)

90,000

Recommended preventive
action limit

Groundwater

Notes:

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a The information was compiled from two sources: 1) EPA regional office outreach by EPA's Office of Science and Technology
(OST) in March 2022; and 2) information from the Interstate Technology and Regulatory Council's (ITRC) Standards and
guidance values forPFAS in groundwater, drinking water, and surface water/effluent (wastewater) PFAS Water and Soil
Values Table, last updated in April 2022 (available for download here: https://pfas-1. .itrcweb.org/fact-slieets/).
b Only states with final guidelines are included; other states may be developing guidelines for PFBS.

In 2020, the European Chemicals Agency (ECHA) adopted an agreement that identified PFBS as
a "Substance of Very High Concern" (ECHA, 2020) based on a "very high potential for
irreversible" human and environmental health effects, and properties including moderate
bioaccumulation in humans, high persistence and mobility in the environment, high potential for
long-range transport, and difficulty of remediating and purifying water.

Table 2 provides drinking water guideline values for PFBS that were developed by international
agencies. The international guideline values range from 90 to 15,000 ppt or ng/L.

Table 2. International Guideline Values for PFBS

Country :l h

PFBS Level
(ppt |ng/L|)

Standard/Guidance

Type of Medium

Reference

Canada

15,000

Screening value

Drinking water

Health Canada (2016)

European
Union (EU)

100 ng/Lc d

Parametric value

Water intended for
human consumption

EU (2020)

500 ng/Lce

Parametric value

Water intended for
human consumption

Denmark

100f

Health based

Groundwater

Danish EPA (2021)

Germany

6,000

Significance
threshold

Groundwater

Von der Trenck et al.
(2018)

Italy

3,000

Environmental
quality standard

Drinking water

Valsecchi et al. (2017)

Sweden

90s

Administrative

Drinking water

Concawe (2016)

Notes:

a The information was collected from the Interstate Technology and Regulatory Council's (ITRC) Standards and guidance values
forPFAS in groundwater, drinking water, and surface water/effluent (wastewater) PFAS Water and Soil Values Table, last
updated in April 2022 (available for download here: https://pfas-1. .itrcweb.org/fact-slieets/).
b Only countries with guideline values provided in the ITRC table are included; other countries may be developing guidelines for
PFBS.

c Parametric values from Directive (EU) 2020/2184 of the European Parliament and of the Council of 16 December 2020 on the
quality of water intended for human consumption. By January 12, 2026, Member States shall take measures necessary to ensure
that water intended for human consumption complies with the parametric values set out in Part B of Annex I in the EU
Directive 2020/2184 (EU, 2020).
d Pertains to a sum of a subset of 20 individual PFAS that includes PFBS: perfluorobutanoic acid (PFBA), perfluoropentanoic
acid (PFPA), perfluorohexanoic acid (PFUxA), perfluoroheptanoic acid (PFHpA), perfluorooctanoic acid (PFOA),
perfluorononanoic acid (PFNA), perfluorodecanoic acid (PFDA), perfluoroundecanoic acid (PFUnDA), perfluorododecanoic
acid (PFDoDA), perfluorotridecanoic acid (PFTrDA), PFBS, perfluoropentane sulfonic acid (PFPS), perfluorohexane sulfonic
acid (PFHxS), perfluoroheptane sulfonic acid (PFHpS), perfluorooctane sulfonic acid (PFOS), perfluorononane sulfonic acid
(PFNS), perfluorodecane sulfonic acid (PFDS), perfluoroundecane sulfonate (PFUnS), perfluorododecane sulfonic acid
(PFDoS), perfluorotridecane sulfonic acid (PFTrS)
e Total PFAS

f Applies to the individual results for PFOA, PFOS, PFNA, PFBA, PFBS, PFHxS, PFHxA, PFHpA, perfluorooctanesulfonamide
(PFOSA), PFDA, 6:2 fluorotelomer sulfonic acid (6:2 FTS), PFPS, PFHpS, PFNS, PFDS, PFUnS, PFDoS, PFTrS, PFPA,
PFUnDA, PFDoDA, PFTrDA as well as the sum of concentrations of these 22 PFAS.
g This limit also applies to the sum of PFOS, PFHxS, PFBS, PFOA, PFHpA, PFHxA and PFPeA.

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1.3	Uses and Sources of PFBS

PFBS is 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; U.S. EPA, 2007). PFBS and its potassium salt were listed on the original
EPA Toxic Substances Control Act (TSCA) Chemical Substance Inventory4 as existing
chemicals that were already in commerce when TSCA was enacted in 1976 (15 U.S.C. § 2601 et
seq.). Therefore, PFBS and its potassium salt were not subject to the pre-manufacture notice
(PMN) reporting process. They are listed as "active" on the inventory but have not been
reviewed under the TSCA New Chemicals program.5 EPA also evaluates existing chemicals
under amended TSCA;6 however, to date, PFBS has not been designated as a high priority
substance for risk evaluation. PFBS and its potassium salt are subject to Section 8 Chemical Data
Reporting.7 While there has not been recent reporting on PFBS, in 2020 there was a report on the
potassium salt (K+PFBS) for one industrial processing and use scenario8 but not for
consumer/commercial uses.

Prior to its use as a replacement chemical, PFBS had been produced solely 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 in
humans of longer-chain PFAS resulted in the use of shorter-chain PFAS such as PFBS as
replacements for longer-chain PFAS in consumer products and applications (U.S. EPA, 2021a).
PFBS and other shorter-chain PFAS possess the desired chemical properties of longer-chain
PFAS, but have shorter half-lives in humans (U.S. EPA, 2021a).

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 (U.S. EPA, 2021a). PFBS has also been used as a mist
suppressant for chrome electroplating and has been detected in association with the use of
aqueous film-forming foam (AFFF) (U.S. EPA, 2021a). PFBS has been detected in dust,
carpeting and carpet cleaners, floor wax, and food packaging (ATSDR, 2021; U.S. EPA, 2021a).

1.4	Environmental Fate, Occurrence in Water, and Exposure to Humans
1.4.1 Environmental Fate and Transport in the Environment

The ionic nature of PFAS, including PFBS, influences physicochemical properties such as water
or lipid solubility and bioaccumulative potential, which impacts environmental fate and transport
and potential human health and ecological effects after exposure (U.S. EPA, 2021a). ECHA
reports that PFBS is stable to hydrolysis, oxidation, and photodegradation in the atmosphere, and

4	TSCA Inventory. Available at https://www.epa.gov/tsca-iiiventory/how-access-tsca-iiiventory

5	Mandated by section 5 of TSCA, EPA's New Chemicals program helps manage the potential risk to human health and the
environment from chemicals new to the marketplace. Section 5 of TSCA is available at https://www.epa. gov/assessing-and-
managing-chemicals-under-tsca/15-usc-ch-53-toxic-substances-control-act

6	On June 22,2016, President Obama signed the Frank R. Lautenberg Chemical Safety for the 21st Century Act, which updates
TSCA. Available at https://www.c0ngress.g0v/l 1.4/plaws/publ1.82/PLAW- 114pufatl82.pdf

7	Basic information about Chemical Data Reporting available here https://www.epa.gov/chemical-data-reporting/basic-
information-about-chemical-data-reporting

8	Section 8 reporting: Processing—incorporation into formulation, mixture, or reaction product; Sector: Electrical Equipment,
Appliance, and Component Manufacturing; Function Category: Flame retardant

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there have been no reports of abiotic degradation under environmental conditions (ECHA, 2019).
PFBS has a high solubility in water (52.6 g/L at 22.5-24 °C for the potassium salt) and high
mobility in the environment (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 (REACH) (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.

1.4.2 Occurrence in Water

PFBS can enter the aquatic environment through releases from manufacturing sites, industrial
uses, fire/crash training areas, and wastewater treatment facilities, as well as from the use of
contaminated biosolids (ATSDR, 2021; U.S. EPA, 2021a). PFBS has been found in rain as well
as in snow/ice in the Arctic and Antarctic (Arp and Slinde, 2018). EPA collected information
about PFBS occurrence in water (described below and in Appendix B, Tables B-l to B-3). To
better understand PFBS sources and occurrence patterns in water, this section includes studies
conducted within and outside the United States. Overall, studies that analyzed water from sites
receiving inputs from or in proximity to known sources of PFAS (as reported by study authors)
did not provide a consistent pattern of detection; increased PFBS detection frequencies (DFs) or
concentrations were not only observed in studies of sites with known sources of PFAS
contamination. Specifically, DFs of 0% were reported at some sites with known, suspected, or
historic PFAS contamination, and DFs of 100% were reported at some sites with no known
sources of PFAS contamination. However, the maximum reported PFBS concentrations were
measured at sites with known PFAS contamination from manufacturing facilities (drinking
water) (Pitter et al., 2020) or AFFF usage (groundwater and surface water) (Anderson et al.,
2016).

1.4.2.1 Drinking Water

EPA required the most nationally representative sampling for PFBS in drinking water to date
under the UCMR 3. Sampling for the UCMR 3 was conducted between 2013 and 2015. PFBS
was detected above the minimum reporting level (MRL)9 of 90 ng/L in eight PWSs (across four
U.S. states and one U.S. territory) out of a total of 4,920 PWSs with results (U.S. EPA, 2017).
PFBS concentrations ranged from 90 (the MRL) to 370 ng/L. Results are available in EPA's
National Contaminant Occurrence Database (NCOD).10 EPA included PFBS among the analytes
that will be monitored under the UCMR 5 and will use EPA analytical Method 533, which was
demonstrated through multilab validation of the method to support a lower UCMR 5 defined
MRL of 3 ng/L.

9	The MRL refers to the quantitation level selected by EPA to ensure reliable and consistent results. It is the minimum
quantitation level that can be achieved with 95 percent confidence by capable analysts at 75 percent or more of the laboratories
using a specified analytical method (EPA, 202 lg).

10	EPA's NCOD is available at https://www.epa.gov/sdwa/national-contamiiiant-occurrence-database-iicod

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Some states have monitored for PFBS in drinking water since the UCMR 3 using improved EPA
analytical methods 533 and 537.1 (see Section 5.0). PFBS has been detected in the finished
drinking water from at least 17 U.S. states (ADEM, 2020; CADDW, 2021; CDPHE, 2020a;
Illinois EPA, 2021b; KYDEP, 2019; MA EE A, 2020; Maine DEP, 2020; MDE, 2021; Michigan
EGLE, 2021; NCDEQ, 2021; NHDES, 2021; NJDEP, 2021; NMED, 2021; Ohio DOH, 2021;
PADEP, 2021; SCDHEC, 2020; VTDEC, 2021). State drinking water PFBS monitoring studies
often focus on investigating areas known to be affected by PFBS. In states where samples were
collected using random sampling site selection (AL, CO, IL, KY, MA, MI, NH, ND, NJ, OH,
SC, and VT), PFBS concentrations ranged from non-detect (ND) to 310 ng/L (ppt). Where
monitoring was targeted to areas known or suspected to have sources of PFBS (CA, ME, MD,
NC, and PA), concentrations were higher and the percentage of samples with PFBS
concentrations above the reporting limit often exceeded 20%. Based on the available finished
drinking water sampling from states, no finished drinking water samples from any state had
PFBS at concentrations exceeding 310 ng/L.

Peer-reviewed studies on PFBS occurrence in drinking water (including bottled water, tap water,
and well water intended for consumption) reporting results from North America and/or Europe
were reviewed (see literature search methods in Appendix A and study details in Appendix B,
Table B-l).

Seven studies analyzed drinking water in areas of North America where study authors did not
indicate whether sampling sites were associated with known or suspected sources of PFAS
release (Appleman et al., 2014; Boone et al., 2014, 2019; Bradley et al., 2020; Dasu et al., 2017;
Hu et al., 2019; Kabore et al., 2018; Subedi et al., 2015). Three of these seven studies (Appleman
et al., 2014; Boone et al., 2019; Bradley et al., 2020) evaluated finished or treated water from
drinking water treatment plants (DWTPs). Appleman et al. (2014) detected PFBS in 100% of
finished water samples taken from DWTPs that used surface water, groundwater, or blended
water as source water, some of which were reportedly known to have been impacted by upstream
wastewater effluent discharge. PFBS levels ranged from 0.43 - 37 ng/L across 11 sites with
finished water samples. Boone et al. (2019) also reported that some sampling locations in their
study had known or suspected sources of wastewater in the source water but did not identify
which ones; PFBS levels in this study ranged from ND to 11.9 ng/L. Bradley et al. (2020)
reported PFBS concentrations of ND-0.5 ng/L in treated pre-distribution tap water from four
sites. Six studies analyzed tap water from homes (Boone et al., 2014; Bradley et al., 2020; Dasu
et al., 2017; Hu et al., 2019; Kabore et al., 2018; Subedi et al., 2015). Across these six studies,
PFBS was detected in at least one sample per study (DFs 5—100%) at concentrations ranging
from ND to 14.15 ng/L; in three of the six studies, the maximum PFBS concentration was < 1
ng/L. In Boone et al. (2014), tap water (for which Mississippi River water was the source) was
tested at one private home during both low and high river stages, and PFBS concentrations were
14.15 ng/L and 2.12 ng/L, respectively. In Hu et al. (2019), the tested water samples were
archived samples from 1989-1990 (PFBS concentrations in these samples ranged from ND-2.97
ng/L).

Three studies conducted in North America examined PFBS levels in drinking water from areas
with known or suspected PFAS releases (Boone et al., 2014; Lindstrom et al., 2011; Scher et al.,
2018) and two of the three studies detected PFBS. Boone et al. (2014) analyzed samples from
three drinking-water wells at sites impacted by AFFF. PFBS was found in all three wells (mean

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PFBS concentrations 9.09-29 ng/L). Lindstrom et al. (2011) sampled six drinking-water wells in
areas impacted by up to 12 years of field applications of biosolids contaminated by a
fluoropolymer manufacturer. PFBS was detected in four of the six wells, and concentrations
were as high as 56.5 ng/L (mean PFBS concentration was 19.7 ng/L). Scher et al. (2018) found
no PFBS in tap water from exterior taps of 23 homes near a former 3M PFAS production facility,
20 of which had been identified as being located within the groundwater contamination area
(GCA).

Of the available studies conducted in Europe, 17 analyzed drinking water samples at sites for
which authors did not indicate whether there were any known associations with PFAS sources or
releases. Fourteen of these 17 studies analyzed tap water from private and/or public sources
(cafes, homes, offices, public fountains); of these 14 studies, 12 detected PFBS in at least one
sample. Across these 12 studies, mean PFBS concentrations ranged from 0.015 in Sweden
(Filipovic and Berger, 2015) to 13.2 ng/L in the Netherlands (Ullah et al., 2011) and the
maximum PFBS concentration was 69.43 ng/L (Barcelona; Ericson et al., 2009). Four of the 17
studies (Boiteux et al., 2012; Eriksson et al., 2013; Eschauzier et al., 2012, 2013) analyzed
finished or treated water at DWTPs, and PFBS levels in these studies ranged from ND in the
Faroe Islands (Eriksson et al., 2013) to 24 ng/L in the Netherlands (Eschauzier et al., 2012).

Nine European studies analyzed drinking water samples from areas near fluoropolymer
manufacturing facilities, AFFF-contaminated military airfields, or fire training sites that may use
AFFF. Six of the nine studies detected PFBS, with maximum concentrations ranging from 11 to
765 ng/L (Brandsma et al., 2019; Gebbink et al., 2017; Gyllenhammar et al., 2015; Li et al.,
2018; Pitter et al., 2020; Weiss et al., 2012). The other three studies (all performed in France)
found no detectable levels of PFBS in treated water from DWTPs located downstream of
fluorochemical manufacturing facilities or a wastewater treatment plant (WWTP) that processes
raw sewage from a fluorochemical manufacturing facility (Bach et al., 2017; Boiteux et al.,
2017; Dauchy et al., 2012). Among the six studies that detected PFBS, the highest measured
PFBS concentration (765 ng/L) was detected in municipal water in Veneto, Italy, sampled from
areas near a fluoropolymer manufacturing facility (Pitter et al., 2020). The study authors reported
that the facility was the only likely source of PFAS and estimated a groundwater contamination
plume with an area of 190 square kilometers (km2) affecting public and private drinking water
sources (Pitter et al., 2020). In the studies that analyzed water samples from areas near AFFF-
contaminated military airfields or fire training sites (Gyllenhammar et al., 2015; Li et al., 2018;
Weiss et al., 2012), PFBS DFs ranged from 0 to 100%, PFBS concentrations ranged from ND to
130 ng/L, and maximum PFBS concentrations ranged from 11 to 130 ng/L.

1.4.2.2 Bottled Water

The United States does not have standards for PFAS in bottled water. The Standard of Quality
set by the International Bottled Water Association (IBWA) for PFAS in bottled water is 5 ng/L
for one PFAS and 10 ng/L for more than one PFAS (IBWA, 2022). One available study analyzed
bottled water in the United States (101 samples representing 66 brands) and reported a PFBS DF
of 17% and PFBS concentrations ranging from ND to 1.44 ng/L (Chow et al., 2021). Of eight
available studies that analyzed bottled water in Canada (one study) or Europe (seven studies), the
study in Canada detected PFBS in 9% of samples at a maximum PFBS concentration of 0.23
ng/L (Kabore et al., 2018). Four of seven studies that analyzed bottled water in different
European countries detected PFBS at concentrations ranging from ND to 51 ng/L (DF 0-29%);

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however, most of the studies did not specify the origin of the bottled water (Gellrich et al., 2013;
Harrad et al., 2019; Le Coadou et al., 2017; Unlii Endirlik et al., 2019). The other three European
studies did not detect PFBS in bottled water.

1.4.2.5 Groundwater

In addition to the studies described in Section 1.4.2.1 that reported groundwater PFBS
concentrations in well water intended for direct consumption, several other studies evaluated the
occurrence of PFBS in raw groundwater in the United States or Europe (see Table B-2). Most of
the available studies sampled from groundwaters known or suspected to be contaminated with
PFAS through various sources, as reported by the study authors. Importantly, some of these
groundwaters are known to be used as input sources for PWSs.

Four U.S. studies assessed PFBS concentrations in groundwater at sites known to be
contaminated with PFAS from the use of AFFF (Anderson et al., 2016; Eberle et al., 2017;
Moody et al., 2003; Steele et al., 2018). Of the three studies that reported PFBS detections, two
reported DFs of 78.26% and 100% (Anderson et al., 2016; Eberle et al., 2017); the third study
did not report a PFBS DF across sample sites but indicated a range of PFBS concentrations (ND-
48 ng/L) (Steele et al., 2018). The fourth study, which analyzed groundwater from the
decommissioned Wurtsmith Air Force Base, did not detect PFBS at any of the ten sites sampled,
though other PFAS were detected (Moody et al., 2003). However, a case study published by the
Association of State and Territorial Solid Waste Management Officials (ASTSWMO) reported
quantifiable levels of PFBS in four of seven samples tested from the Wurtsmith Air Force Base;
one site sampled directly below the fire training area was reported to have a PFBS concentration
of 4,100 ng/L (ASTSWMO, 2015).

Additionally, PFBS has been detected at concentrations ranging from 0.00211 ng/L to 0.0261
ng/L in groundwater wells (100% well DF) at a site near the 3M Cottage Grove
perfluorochemical manufacturing facility in Minnesota (3M, 2007; ATSDR, 2021). Lee et al.
(2015) evaluated urban shallow groundwater contaminated by wastewater effluent discharge and
reported a DF of 20% (1 of 5 shallow sites) and a maximum PFBS level of 36.3 ng/L. In
contrast, Procopio et al. (2017) collected groundwater from 17 sampling sites (53 total across all
water types sampled), some of which were located downstream of an industrial facility that used
materials containing PFOA. PFBS was not detected in groundwater collected from any of the
sampling locations. Post et al. (2013) assessed raw water from PWS intakes in New Jersey; these
intake locations were selected to represent New Jersey geographically and they were not
necessarily associated with any known PFAS release. PFBS was detected pre-treatment in 1 of
18 systems at a concentration of 6 ng/L (MRL = 5 ng/L). Lindstrom et al. (2011) analyzed water
from 13 wells intended for uses other than drinking water (e.g., livestock, watering gardens) in
areas impacted by up to 12 years of field applications of biosolids contaminated by a
fluoropolymer manufacturer. PFBS was detected in three of the wells (mean concentration 10.3
ng/L; range: ND-76.6 ng/L).

Of the 10 identified studies conducted in Europe, seven studies evaluated groundwater samples
from sites with known or suspected PFAS releases associated with AFFF use, fluorochemical
manufacturing, or other potential emission sources including landfill/waste disposal sites, skiing
areas, or areas of unspecific industries that use PFAS in manufacturing (e.g., metal plating)
(Dauchy et al., 2012, 2017, 2019; Gobelius et al., 2018; Gyllenhammar et al., 2015; Fteisseter et

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al., 2019; Wagner et al., 2013). All of these studies reported PFBS detections in at least one
sample or site, though only two studies (both conducted in the vicinity of areas with known
AFFF usage) reported PFBS concentrations >100 ng/L (Dauchy et al., 2019; Gyllenhammar et
al., 2015). The remaining three studies of the 10 identified did not provide information on
whether there were potential sources of PFAS at the sampling locations or were designed to be
regionally, nationally, or internationally representative (Barreca et al., 2020; Boiteux et al., 2012;
Loos et al., 2010). At these sites, PFBS was detected infrequently (DFs 4 to 18%) with a
maximum concentration of 25 ng/L across the three studies.

1.4.2.4 Surface Water

Studies evaluating the occurrence of PFBS in surface water are available from North America,
Europe, and across multiple continents (see Table B-3). 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 ND to 6.2 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 a range of PFBS concentrations from ND - 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 DF in the Nakayama et al.
(2010) study was 43% with median and maximum PFBS levels of 0.71 and 84.1 ng/L,
respectively. Pan et al. (2018) sampled surface water sites in the Delaware River and reported a
100%) DF, though PFBS levels were relatively low (0.52 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 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 WWTPs (Galloway et al.,
2020; Zhang et al., 2016). Among these sites, DFs (0 to 100%) and PFBS levels (ND to 336
ng/L) varied. In general, DFs that ranged from 0 to 3% were associated with samples collected
upstream of PFAS point sources, and higher 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

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applications of biosolids contaminated by a fluoropolymer manufacturer, and the maximum and
mean PFBS concentrations were 208 and 26.3 ng/L, respectively.

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 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 higher 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.

Seven studies evaluated 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, DFs varied (0 to 100%) and PFBS levels were < 250 ng/L overall.

Ten studies conducted in Europe evaluated sites with no known fluorochemical source of
contamination (Ahrens et al., 2009a, 2009b; Barreca et al., 2020; Ericson et al., 2008b; 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). 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 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 |ig/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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1.4.3 Exposure in Humans

As described in EPA's final PFBS toxicity assessment, PFBS has been detected in the serum of
humans in the general population (U.S. EPA, 2021a). In American Red Cross plasma samples
collected in 2015, 8.4% of samples had a quantifiable serum PFBS concentration, ranging from
the lower limit of quantitation (LOQ) to 4.2 nanograms per milliliter (ng/mL) (Olsen et al.,
2017). Results for the majority of serum samples were below the lower LOQ for PFBS, and the
95th percentile concentration was 0.02 ng/mL (Olsen et al., 2017). Data from the 2013-2014
National Health and Nutrition Examination Survey (NHANES) reported a 95th percentile
concentration for PFBS in serum that was at or below the level of detection (0.1 ng/mL) (Olsen
et al., 2017). Another study studied temporal trends of PFBS in blood serum from primiparous
nursing women in Sweden -2000-2002 around the time of increased manufacturing of PFBS
after it was introduced as a replacement for PFOS (Glynn et al., 2012). An increase in PFBS
blood serum levels was observed between 1996 and 2010, and regression analysis suggested that
PFBS levels doubled on average every six years (Glynn et al., 2012).

Studies in animals show that PFBS is well absorbed following oral administration and distributes
to all tissues of the body (Bogdanska et al., 2014). Distribution is predominantly extracellular
(Olsen et al., 2009) and based on its resistance to metabolic degradation, the majority of PFBS is
eliminated unchanged in urine and feces. Two studies that measured PFBS half-life in humans
found overlapping ranges of 21.6-87.2 days (Xu et al., 2020) and 13.1-45.7 days (Olsen et al.,
2009). The relatively rapid rate of elimination (days to weeks) of PFBS, compared with longer-
chain PFAS (years), could lead to a lack of detection in biomonitoring detects which should not
be interpreted as a lack of occurrence or exposure potential (U.S. EPA, 2021a). For more
information, see U.S. EPA (2021a).

2 J Problem Formulation and Scope
2.1 Conceptual Model

A conceptual model provides useful information to characterize and communicate the potential
health risks related to PFBS exposure from drinking water and to outline the scope of the HA.
The sources of PFBS, the routes of exposure for biological receptors of concern (e.g., various
human activities related to tap water ingestion such as drinking, food preparation, and
consumption), the potential health effects, and exposed populations including sensitive
populations and life stages are depicted in the conceptual diagram below (Figure 1).

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STRESSOR(S)

PFBS and its Potassium Salt

POTENTIAL
SOURCES



Ambient

















Drinking
Water

Ground

and
Surface
Water

Industrial

Uses

Air

SoU

Food

Dust

Consumer
Products

Fire-
Fighting
Foams

Biosolids

RSC Derivation

POTENTIAL
EXPOSURE
ROUTES

Final Toxicity
Assessment for
PFBS
(U.S. EPA, 2021a)

Oral





Inhalation

(includes incidental inhalation during
showering'bathing)

(includes drinking water, cooking with
water, incidental ingestion during
showering bathing)

Dermal

(includes showering bathing)

AFFECTED
HEALTH
OUTCOMES

POTENTLVLLY SENSITIVE
POPULATIONS WITHIN
GENERAL POPULATION

Adults

C hildren (including

breastfed and/or
formula-fed infants)

W omen of Child-
Bearing Age

Pregnant Women
and Developing
Embryo s/F etuses

Lactating
Women

Figure 1. Conceptual Model for PFBS Drinking Water Health Advisory Development

EF Selection

Legend

Dark Blue = Information used for
deriving HA value

Light Blue = Information available
but not used for deriving HA value

Gray = Out of scope for deriving HA
value

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The conceptual model is intended to explore potential links between exposure to a contaminant
or stressor and the adverse health outcomes, and to outline the information sources used to
identify or derive the input values used for the HA derivation, which are the RfD, relative source
contribution (RSC), and exposure factor (EF). The conceptual model also illustrates the scope of
the PFBS HA, which considers the following factors:

Stressors: The scope of this drinking water HA includes PFBS, its potassium salt (K+PFBS), and
PFBS- since K+PFBS fully dissociates in water to the deprotonated anionic form of PFBS
(PFBS-; CASRN 45187-15-3) and the K+ cation at environmental pH levels (pH 4-9), consistent
with the scope of the PFBS toxicity assessment (EPA, 2021a).

Potential Sources of Exposure: The scope of the HA derivation is limited to drinking water
from public water facilities or private wells. Sources of PFBS exposure include both ground and
surface waters used for drinking. To develop the RSC, information about non-drinking water
sources was identified to determine the portion of the RfD attributable to drinking water.

Potential non-drinking water sources of PFBS include but are not limited to foods, indoor dust,
indoor and outdoor air, soil, biosolids, and consumer products (see Figure 1).

Potential Exposure Routes: Oral exposure to PFBS from contaminated drinking water sources
(e.g., via drinking water, cooking with water, and incidental ingestion from showering) is the
focus of the HA. The drinking water HA value does not apply to other exposure routes.

However, information on other potential routes of exposure including dermal exposure (contact
of exposed parts of the body with water containing PFBS during bathing or showering,
dishwashing); and inhalation exposure (during bathing or showering or using a humidifier or
vaporizer) was considered to develop the RSC.

Affected Health Outcomes: The PFBS final toxicity assessment (U.S. EPA, 2021a) considered
all publicly available human, animal, and mechanistic studies of PFBS exposure and effects. The
assessment identified associations between PFBS exposure and thyroid, developmental, and
kidney effects. As part of the PFBS final toxicity assessment, human and animal studies of other
health effects after PFBS exposure included the evaluation of effects on the reproductive system,
liver, and lipid and lipoprotein homeostasis but the evidence did not support clear associations
between exposure and effect. No cancer studies were identified for PFBS (U.S. EPA, 2021a).

Potentially Sensitive Populations or Life Stages: The receptors are humans in the general
population who could be exposed to PFBS from oral exposure to tap water through ingestion at
their homes, workplaces, schools, and daycare centers. Within all ages of the general population,
there are potentially sensitive populations or life stages that may be more susceptible due to
increased exposure and/or response. Potentially sensitive populations include the developing
embryo and fetus (exposed to PFBS via the pregnant woman) and women of childbearing age
who may be or become pregnant.

2.2 Analysis Plan

2.2.1 Health Advisory Guidelines

Assessment endpoints for HA guidelines or values can be developed, depending on the available
data, for both short-term (one-day and ten-day) and lifetime exposure using information on the
noncarcinogenic and carcinogenic toxicological endpoints of concern. Where data are available,

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HAs can reflect sensitive populations or life stages that may be more susceptible and/or more
highly exposed.

One-Day HA is protective of noncancer effects for up to 1 day of exposure and is
typically based on an in vivo toxicity study with a duration of 7 days or less. It is
typically calculated for an infant.

Ten-Day HA is protective of noncancer effects for up to 10 days of exposure and is
typically based on an in vivo toxicity study with a duration of 7 to 30 days. It is
typically calculated for an infant.

Lifetime HA is designed to be protective of noncancer effects over a lifetime of
exposure and is typically based on a chronic in vivo experimental animal toxicity
study and/or human epidemiological data.

10"6 Cancer Risk Concentration is the concentration of a carcinogen in water at
which the population is expected to have a one in a million (10"6) excess cancer risk
above background after exposure to the contaminant over a lifetime. It is calculated
for carcinogens classified as known or likely human carcinogens (U.S. EPA, 1986,
2005b). Cancer risk concentrations are not derived for substances for which there is
suggestive evidence of carcinogenic potential unless the cancer risk has been
quantified.

2.2.2 Sources of Toxicity Information for Health Advisory Development
The final toxicity assessment for PFBS, entitled Human Health Toxicity Values for
Perfluorobutane Sulfonic Acid (CASRN 375-73-5) and Related Compound Potassium
Perfluorobutane Sulfonate (CASRN29420-49-3), published in April 2021 by EPA's Office of
Research and Development (ORD) Center for Public Health and Environmental Assessment
(CPHEA) (U.S. EPA, 2021a), serves as the basis of the toxicity information and chronic RfD
used to derive the lifetime noncancer HA for PFBS. It also synthesizes and describes other
information on PFBS including physicochemical properties and toxicokinetics. The PFBS
toxicity assessment was published after rigorous scientific review, including internal and external
review, and public comment.

To develop the final toxicity assessment for PFBS, EPA reviewed and analyzed the available
toxicokinetics and toxicity data for PFBS. Briefly, online scientific databases (PubMed, Web of
Science, TOXLINE, and TSCATS via TOXLINE) were searched using search terms focused on
chemical name and synonyms with no limitations on publication type, evidence stream (i.e.,
human, animal, in vitro, and in silico), or health outcomes. The identified studies were screened
using Populations, Exposures, Comparators, and Outcomes (PECO) criteria and relevant studies
underwent study quality evaluation. Dose-response studies were identified for dose-response
modeling and a point-of-departure (POD) and uncertainty factors (UFs) were selected for RfD
derivation. For more information, please see Section 2.3 in U.S. EPA (2021a).

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2.2.3 Approach and Scope for Health Advisory Derivation
2.2.3.1 Approach for Deri ving Noncancer HAs

The following equations (Eqs. 1-3) are used to derive the HAs.11 Lifetime HAs and 10"6 cancer
risk concentrations are only derived for chemicals without an existing National Primary Drinking
Water Regulation.

( POD \

Oneway HA =(UFc,DW,.BW)

POD is typically derived from a toxicity study of duration 7 days or less

(Eq. 1)

( POD \

Ten-Day HA =(UFc^DWI BW)

POD is typically derived from a toxicity study of duration 7-30 days

(Eq. 2)

/ RfD \

Lifetime HA = I ^T,TT ^T,T) * RSC
VDWI-BW/

RfD is typically derived from a chronic study

(Eq. 3)

Where:

POD is the point of departure, typically a lowest observed adverse effect level (LOAEL), a no
observed adverse effect level (NOAEL), or a BMDL from the critical study.

UFc is the composite UF or total UF value after multiplying individual UFs. UFs are established
in accordance with EPA best practices (U.S. EPA, 2002) and consider uncertainties related to the
following: variation in sensitivity among the members of the human population (i.e., inter-
individual variability), extrapolation from animal data to humans (i.e., interspecies uncertainty),
extrapolation from data obtained in a study with less-than-lifetime exposure to lifetime exposure
(i.e., extrapolating from subchronic to chronic exposure), extrapolation from a LOAEL rather
than from a NOAEL, and extrapolation when the database is incomplete. For PFBS, the value of
UFc was determined in the final PFBS toxicity assessment (U.S. EPA, 2021a).

DWI-BW is the 90th percentile drinking water intake (DWI), adjusted for body weight (bw), for
the selected population in units of liter per kilogram body weight per day (L/kg bw-day). The
DWI-BW considers direct and indirect consumption of tap water (indirect water consumption
encompasses water added in the preparation of foods or beverages, such as tea and coffee). For
PFBS, the value of this parameter is based on the critical study identified in the PFBS final
toxicity assessment (U.S. EPA, 2021a), and is identified in Chapter 3 of EPA's Exposure Factors
Handbook (U.S. EPA, 2019a).

11 https://www.epa.gov/system/files/documents/2022-01/dwtable2018.pdf

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RfD is the reference dose—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. The value of this parameter
was derived in the final PFBS toxicity assessment and is based on the critical effect and study
identified in that assessment (U.S. EPA, 2021a).

RSC is the relative source contribution—the percentage of the total oral exposure attributed to
drinking water sources (U.S. EPA, 2000a) where the remainder of the exposure is allocated to
other routes or sources. The RSC is calculated by examining other sources of exposure (e.g., air,
food, soil) and pathways of exposure in addition to drinking water using the methodology
described for calculation of an RSC described in U.S. EPA (2000a) and Section 3.3.

2.2.3.2	Scope of Noncanc lih Advisory Values

Adequate data are available to derive a lifetime HA for PFBS. Neither one-day nor ten-day HA
values were derived for PFBS. U.S. EPA (2021a) derived subchronic and chronic RfDs but did
not derive an RfD for exposure durations of 7 days or less on which to base a one-day HA for
PFBS. Derivation of a 10-day HA was considered because the subchronic and chronic RfDs are
both based on a 20-day exposure study, which may be used to derive a ten-day HA. However,
the critical health effect on which the chronic RfD used to calculate the lifetime HA is based
(i.e., decreased serum levels of the thyroid hormone thyroxine [T4] in newborn mice) resulted
from PFBS exposure during a developmental life stage. EPA's risk assessment guidelines for
developmental toxicity indicate that adverse effects can result from even brief exposure during a
critical period of development (U.S. EPA, 1991). The critical study for the subchronic and
chronic RfDs for PFBS observed persistent health effects into adulthood suggesting the potential
for long-term health consequences of gestational-only PFBS exposure and that gestation is at
least one critical exposure window for PFBS. Therefore, the lifetime HA (calculated in Section
4.0) and the chronic RfD from which it is derived (see Table 4) are considered applicable to
short-term PFBS exposure scenarios (including during pregnancy) via drinking water.

2.2.3.3	Approach and Scope for Deriving Cancer Risk Concentrations

The following equations (Eqs. 4-5) are used to derive cancer risk concentrations.

Calculated for non-mutagenic carcinogens12 only:

lxlO-6

10 b Cancer Risk Concentration =	

CSF * DWI-BW

(Eq. 4)

Calculated for mutagenic carcinogens only:

lxlO-6 v1 /Fi*ADAFi\

10 b Cancer Risk Concentration = —^..T

CSF A; V DWI-BW; /

(Eq. 5)

12 https://www.epa.gov/system/files/documents/2022-01/dwtable2018.pdf

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Where:

CSF is the cancer slope factor—an upper bound, approximating a 95 percent confidence limit of
the increased cancer risk from a lifetime of oral exposure to a stressor. The value for this
parameter is derived in the final toxicity assessment when data are available.

DWI-BWi is the 90th percentile bw-adjusted DWI in units of L/kg bw-day for each age group
(i), considered when calculating cancer risk concentrations for mutagenic carcinogens.

ADAFi is the age-dependent adjustment factor for each age group (i), used when calculating
cancer risk concentrations for carcinogens that act via a mutagenic mode of action (U.S. EPA,
2005a,b).

Fi the fraction of life spent in each age group (i), used when calculating cancer risk
concentrations for mutagens (U.S. EPA, 2005a).

2.2.3.4 Scope of Cancer Risk Concentration Deri vation

As described in the toxicity assessment for PFBS, a CSF was not derived because no studies
evaluating the carcinogenicity of PFBS in humans or animals had been identified (U.S. EPA,
2021a). In accordance with the Guidelines for Carcinogen Risk Assessment (U.S. EPA, 2005b),
EPA concluded that there is "Inadequate Information to Assess Carcinogenic Potential' for
PFBS by any route of exposure (U.S. EPA, 2021a). Therefore, a 10"6 cancer risk concentration
cannot be derived for PFBS at this time.

2.2.4 Exposure Factors for Deriving Health A dvisory

2.2.4.1 Exposure Factor Selection

An EF, such as body weight-adjusted drinking water intake (DWI-BW), is one of the input
values for deriving a drinking water HA. EFs are factors related to human activity patterns,
behavior, and characteristics that help determine an individual's exposure to a contaminant.
EPA's Exposure Factors Handbook (EFH)13 is a resource for conducting exposure assessments
and provides EFs based on information from publicly available, peer-reviewed studies. Chapter 3
of the EFH presents EFs in the form of DWI and DWI-BW for various populations or life stages
within the general population (U.S. EPA, 2019a). The use of EFs in HA calculations is intended
to protect sensitive populations and life stages within the general population from adverse effects
resulting from exposure to a contaminant.

When developing HAs, the goal is to protect all ages of the general population including
potentially sensitive populations or life stages such as children. The approach to select the EF for
the drinking water HA includes a step to identify potentially sensitive population(s) or life
stage(s) (i.e., populations or life stages that may be more susceptible or sensitive to a chemical
exposure) by considering the available data for the contaminant. Although data gaps can prevent
identification of the most sensitive population (e.g., not all windows of exposure or health
outcomes have been assessed for PFBS), the critical effect and POD that form the basis for the
RfD can provide some information about sensitive populations because the critical effect is
typically observed at the lowest tested dose among the available data. Evaluation of the critical
study, including the exposure interval, may identify a particularly sensitive population or life

13 EPA's EFH is available at https://www.epa.gov/expobox/about-exposure-factors-handbook

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stage (e.g., pregnant women, formula-fed infants, lactating women). In those cases, EPA can
select the corresponding DWI-BW for that sensitive population or life stage from the EFH (U.S.
EPA, 2019a) for use in HA derivation. When multiple potentially sensitive populations or life
stages are identified based on the critical effect or other health effects data (from animal or
human studies), EPA selects the population or life stage with the greatest DWI-BW because it is
the most health protective. For deriving lifetime HAs, the RSC corresponding to the selected
sensitive life stage is also determined when data are available (see Section 3.3). In the absence of
information indicating a potentially sensitive population or life stage, the EF corresponding to all
ages of the general population may be selected.

To derive chronic HAs, EPA typically uses DWI EFs normalized to body weight (i.e., DWI-BW
in liter [L] of water consumed/kg bw-day) for all ages of the general population or for a sensitive
population or life stage, when identified. The Joint Institute for Food Safety and Applied
Nutrition's Food Commodity Intake Database (FCID) Consumption Calculator Tool14includes
the EPA EFs and can also be used to estimate DWI-BW for specific populations, life stages, or
age ranges. EPA uses the 90th percentile DWI-BW to ensure that the HA is protective of the
general population as well as sensitive populations or life stages (U.S. EPA, 2000a, 2016a). In
2019, EPA updated its EFs for DWI-BW based on newly available science (EPA, 2019a).

Table 3 shows EPA EFs that have been derived based on the available data for some sensitive
populations or life stages that are considered when deriving HAs. Other groups may be
considered depending on the available information regarding sensitivity to health effects after
exposure to a contaminant.

Table 3. EPA Exposure Factors for Drinking Water Intake

Populations or
Life Stages

DWI-BW

(L/kg bw-day)

Description of Exposure Metric

Source

General
population, all
ages

0.0338

90th percentile direct and indirect
consumption of community water,
consumer-only two-day average, all
ages.

2019 Exposure Factors
Handbook Chapter 3,
Table 3-21, NHANES
2005-2010 (U.S. EPA,
2019a)

Children

0.143

90th percentile direct and indirect
consumption of community water,
consumer-only two-day average, birth
to < 1 year.

2019 Exposure Factors
Handbook Chapter 3,
Table 3-21, NHANES
2005-2010 (U.S. EPA,
2019a)

Formula-fed
infants

0.249

90th percentile direct and indirect
consumption of community water,
formula-consumers only, 1 to < 3
months. Includes water used to
reconstitute formula, plus all other
community water ingested.

Kahn et al. (2013)
Estimates ofWater
Ingestion in Formula by
Infants and Children
Based on CSFII 1994-
1996 and 1998ab

14 Joint Institute for Food Safety and Applied Nutrition's Food Commodity Intake Database, Commodity Consumption
Calculator is available at https://fcid.foodrisk.org/percentiles

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Populations or
Life Stages

DWI-BW

(L/kg bw-day)

Description of Exposure Metric

Source

Pregnant women

0.0333

90th percentile direct and indirect
consumption of community water,
consumer-only two-day average.

2019 Exposure Factors
Handbook Chapter 3,
Table 3-63, NHANES
2005-2010 (U.S. EPA,
2019a)

Women of
childbearing age

0.0354

90th percentile direct and indirect
consumption of community water,
consumer-only two-day average, 13 to
<50 years.

2019 Exposure Factors
Handbook Chapter 3,
Table 3-63, NHANES
2005-2010 (U.S. EPA,
2019a)

Lactating women

0.0469

90th percentile direct and indirect
consumption of community water,
consumer-only two-day average.

2019 Exposure Factors
Handbook Chapter 3,
Table 3-63, NHANES
2005-2010c (U.S. EPA,
2019a)

Notes: CSFII = continuing survey of food intake by individuals; L/kg bw-day = liter per kilogram body weight per day.
a The sample size does not meet the minimum reporting requirements as described in the Third Report on Nutrition Monitoring in

the United States (LSRO, 1995).
b Chapter 3.2.3 in U.S. EPA (2019a) cites Kahn et al. (2013) as the source of drinking water ingestion rates for formula-fed
infants. While U.S. EPA (2019a) provides the 95th percentile total direct and indirect water intake values, Office of
Water/Office of Science and Technology (OW/OST) policy is to utilize the 90th percentile DWI-BW. OW/OST was able to
identify the 90th percentile DWI-BW in Kahn et al. (2013) and report the value in this table.
c Estimates are less statistically reliable based on guidance published in the Joint Policy on Variance Estimation and Statistical
Reporting Standards on NHANES III and CSFII Reports: Human Nutrition Information Service (HNIS)/National Center for
Health Statistics (NCHS) Analytical Working Group Recommendations (NCHS, 1993).

2.2.4.2 Determining Proportion of RfD A tlribiituble to Drinking Water

To account for aggregate risk from exposures and exposure pathways other than oral ingestion of
drinking water, EPA applies an RSC when calculating HAs to ensure that total human exposure
to a contaminant does not exceed the daily exposure associated with the RfD. The RSC
represents the proportion of an individual's total exposure to a contaminant that is attributed to
drinking water ingestion (directly or indirectly in beverages like coffee, tea, or soup, as well as
from transfer to dietary items prepared with drinking water) relative to other exposure pathways.
The remainder of the exposure equal to the RfD is allocated to other potential exposure sources
(U.S. EPA, 2000a). The purpose of the RSC is to ensure that the level of a contaminant (e.g., HA
value), when combined with other identified sources of exposure common to the population of
concern, will not result in exposures that exceed the RfD (U.S. EPA, 2000a).

To determine the RSC, EPA follows the Exposure Decision Tree for Defining Proposed RfD (or
POD/UF) Apportionment in EPA's guidance, Methodology for Deriving Ambient Water Quality
Criteria for the Protection of Human Health (U.S. EPA, 2000a). EPA considers whether there
are significant known or potential uses/sources other than drinking water, the adequacy of data
and strength of evidence available for each relevant exposure medium and pathway, and whether
adequate information on each source is available to quantitatively characterize the exposure
profile. The RSC is developed to reflect the exposure to the general population or a sensitive
population within the general population exposure.

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Per EPA's guidance, in the absence of adequate data to quantitatively characterize exposure to a
contaminant, EPA typically recommends an RSC of 20%. When scientific data demonstrating
that sources and routes of exposure other than drinking water are not anticipated for a specific
pollutant, the RSC can be raised as high as 80% based on the available data, thereby allocating
the remaining 20% to other potential exposure sources (U.S. EPA, 2000a).

To inform the RSC determination, available information on all exposure sources and routes for
PFBS was identified using the literature search and screening method described in Appendix A.
To identify information on PFBS exposure routes and sources to inform RSC determination,
EPA considered primary literature published between 2003-2020 and collected by EPA ORD as
part of an effort to evaluate evidence for pathways of human exposure to eight PFAS, including
PFBS. In order to consider more recently published information on PFBS exposure, EPA
incorporated the results of a date-unlimited gray literature search that was conducted in February
2022 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.

J J Health Advisory Input Tallies
3.1 Toxicity Assessment Values

Table 4 summarizes the peer-reviewed chronic noncancer toxicity values from EPA's Human
Health Toxicity Values for Perjluorobutane Sulfonic Acid (CASRN 375-73-5) and Related
CompoundPotassium Perfluorobutane Sulfonate (CASRN29420-49-3) (U.S. EPA, 2021a).

Table 4. Chronic Noncancer Toxicity Information for PFBS for Deriving the Lifetime HA

Health Assessment

PFBS
Exposure in
Critical Study

RfD

(mg/kg bw-day)

Critical Effect

Principal Study

Human Health Toxicity
Values for Perfluorobutane
Sulfonic Acid (CASRN 375-
73-5) and Related
Compound Potassium
Perfluorobutane Sulfonate
(CASRN 29420-49-3)

Days 1-20 of
gestation

K+PFBS: 3 x 10"4
PFBS: 3 x lO"4

Decreased
serum total T4
in newborn
(PND 1) mice

Oral gestational
exposure study in
mice (Feng et al.,
2017)

Notes: mg/kg bw-day = milligram per kilogram body weight per day; PND = post-natal day.
Source: U.S. EPA, 2021a

As stated in U.S. EPA (2021a), the thyroid effect of decreased thyroid hormones, specifically
serum total T4, in newborn (PND1) mice exposed to K+PFBS throughout gestation was selected
as the critical effect (Feng et al., 2017). This critical effect and study were used to derive the
chronic RfDs for K+PFBS and PFBS of 3 x 10 4 milligrams per kilogram body weight per day
(mg/kg bw-day).

Based on EPA's Recommended Use of Body Weight4 as the Default Method in Derivation of the
Oral Reference Dose (U.S. EPA, 2011), serum half-lives were used to scale a toxicologically
equivalent dose of orally administered K+PFBS from animals to humans. Following EPA's

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Benchmark Dose Technical Guidance Document (U.S. EPA, 2012b), benchmark dose (BMD)
modeling of thyroid effects following gestational exposure to K+PFBS resulted in a benchmark
dose lower confidence limit for 0.5 standard deviation change from the control (BMDLo.ssd)
human equivalent dose (HED) of 0.095 mg/kg bw-day.

This POD (HED) served as the critical effect and was divided by a composite UF (UFc) of 300.
The UFc is based on an animal-to-human UF (UFa) of 3 to account for extrapolation from mice
to humans; an intrahuman UF (UFh) of 10 to account for interindividual differences in human
susceptibility; and a database UF (UFd) of 10 to account for deficiencies in the toxicity database.
A value of 1 was applied for the extrapolation from subchronic to a chronic exposure duration UF
(UFs) because extrapolation from subchronic to chronic was not needed, and UFl because a
LOAEL to NOAEL approach was not used. 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 mg/kg bw-day), after
adjusting for differences in molecular weight (MW) between K+PFBS (338.19) and PFBS
(300.10) (see Section 6.0 in U.S. EPA [2021a] for more details). This chronic RfD for PFBS was
used to derive the lifetime HA.

No studies evaluating the carcinogenicity of PFBS in humans or animals were identified (U.S.
EPA, 2021a). In accordance with EPA's Guidelines for Carcinogen Risk Assessment (U.S. EPA,
2005b), EPA concluded that there is "Inadequate Information to Assess Carcinogenic Potential"
for PFBS by any route of exposure (U.S. EPA, 2021a) and did not derive a 10"6 cancer risk
concentration.

3.2 Expo sure F actors

To identify potentially sensitive populations or life stages, EPA considered the PFBS exposure
interval used in the critical study selected for chronic RfD derivation in the final PFBS toxicity
assessment (U.S. EPA, 2021a). In the critical study pregnant mice were orally exposed to
K+PFBS throughout all of gestation (days 1-20 of gestation) (Feng et al., 2017; U.S. EPA,
2021a), identifying the developing fetus (exposed via the pregnant mother) as a population that
may be particularly susceptible to PFBS exposure. The critical study did not permit a more
precise identification of the most sensitive or critical PFBS exposure window during prenatal
development since exposure was throughout all of gestation. The critical effect of thyroid
development in the developing mouse embryo and fetus is relevant to humans. Human thyroid
development occurs in three phases during gestation, and while there are some timing differences
in thyroid development between humans and rodents (see Section 6.1.1.3 in U.S. EPA, 2021a),
two phases of thyroid development occur during gestation in both the mouse and human.

The gestational exposure in the critical study is relevant to two potentially sensitive populations
or life stages—women of childbearing age (13 to < 50 years) who may be or become pregnant,
and pregnant women and their developing embryo and fetus (Table 5). EPA selected women of
childbearing age as the sensitive life stage for HA derivation because the DWI-BW is greater
(0.0354 L/kg bw-day) than for pregnant women (0.0333 L/kg bw-day). EPA addresses exposure
to the sensitive developing embryo and fetus because they are exposed to drinking water via the
pregnant mother. Additional support for the women of childbearing age population including
pregnant women (and their developing embryo and fetus) includes the high rate of unintended
pregnancies reported in the United States (30.6%) (United Health Foundation, 2021). To derive
the HA value, EPA used the DWI-BW of 0.0354 L/kg bw-day representing the consumers-only

21


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two-day average of direct and indirect community water consumption at the 90th percentile for
women of childbearing age (13 to < 50 years) (Table 5, in bold).

Table 5. EPA Exposure Factors for Drinking Water Intake for Different Candidate
Sensitive Populations or Life Stages Based on the Critical Effect and Study

Population

DWI-BW

(L/kg bw-day)

Description of Exposure
Metric

Source

Women of
childbearing age

0.0354

90th percentile direct and
indirect consumption of
community water, consumer-
only two-day average, 13 to <
50 years.

2019 Exposure Factors
Handbook Chapter 3,
Table 3-63, NHANES
2005-2010 (U.S. EPA,
2019a)

Pregnant women

0.0333

90th percentile direct and
indirect consumption of
community water, consumer-
only two-day average.

2019 Exposure Factors
Handbook Chapter 3,
Table 3-63, NHANES
2005-2010 (U.S. EPA,
2019a)

Notes'. L/kg bw-day = liters of water consumed per kilogram body weight per day. The DWI-BW used to calculate the PFBS
lifetime HA is in bold.

3.3 Relative Source Contribution

As stated in the analysis plan, EPA collected and evaluated information about PFBS exposure
routes and sources to inform RSC determination. Results from the literature search are described
below.

3.3.1 Non-Drinking Water Sources and Routes

EPA presents information below from studies performed in the United States as well as studies
published globally for this emerging contaminant to be as comprehensive as possible, given that
the overall information is limited. While the studies from non-U.S. countries inform an
understanding global exposure sources and trends, the RSC determination is based on the
available data for the United States.

3.3.1.1 Dietary Sources
Food

PFBS was included in a suite of individual PFAS selected as part of PF AS-targeted
reexaminations of samples collected for the U.S. Food and Drug Administration's (FDA's) Total
Diet Study (U.S. FDA, 2020a,b, 2021 a,b, 2022a,b); however, it was not detected in any of the
food samples tested. It should be noted that 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 (U.S. FDA, 2022c). 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 (U.S. FDA 2018, 2021c). Maximum residue
levels for PFBS were not found in the Global MRL Database (Bryant Christie Inc., 2022).

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In addition to efforts by 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 B-4). 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 B-4)—one in the United
States (von Ehrenstein et al., 2009) and seven in Europe (Abdallah et al., 2020; Beser et al.,
2019; Cariou et al., 2015; Karrman et al., 2007, 2010; Lankova et al., 2013; Nyberg et al., 2018).

Of eight 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 GCA impacted by a former 3M PFAS production facility (PFBS
concentrations ranged from ND to 0.065 nanograms per gram [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.

Based on the available data to date, seafood (including fish and shellfish) has been found to
contain the highest concentrations of PFBS out of all food types examined. Several large-scale
sampling efforts have been conducted by EPA and other agencies to determine PFAS levels in
fish. In EPA's 2013-2014 National Rivers and Streams Assessment (NRSA), PFBS was detected
at concentrations between the quantitation limit (1 ng/g) and the method detection limit (0.1
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 (U.S. EPA, 2020a). Notably, PFBS was not detected in any fish species sampled in
the 2008-2009 NRSA (Stahl et al., 2014). PFBS was also detected at a concentration of 0.36
ng/g in a smallmouth bass fillet composite collected from Lake Erie, New York in EPA's 2015
Great Lakes Human Health Fish Fillet Tissue Study (U.S. EPA, 2021d). PFBS has been detected
in 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, 2022). PFBS was not a target chemical
in EPA's National Lake Fish Tissue Study (U.S. EPA, 2009a).

Several peer-reviewed publications that examined PFBS concentrations in fish and shellfish are
also available. As mentioned previously, Schecter et al. (2010) detected PFBS in cod samples.
Mean PFBS levels in cod from this study (0.12 ng/g wet weight [ww]) were much lower than
maximum levels detected in Alaska blackfish obtained from the Suqi River, Alaska in remote

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locations upstream and downstream of a former (unnamed) defense site (59.2 ng/g ww) (Byrne
et al., 2017). In this study, blackfish were considered sentinel species but are not among the
traditional fish consumed in the area. The authors noted that the presence of PFAS in fish from
remote sites is suggestive of atmospheric deposition. 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).

The European Food Safety Authority (EFSA) 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 nanograms per
kilogram body weight per day (ng/kg bw-day) (minimum) to 0.10-3.72 ng/kg bw-day
(maximum) (EFSA, 2012). Of 27 studies conducted in Europe, 12 found PFBS in at least one
food type (Table B-4). Eight of the 12 studies included food samples obtained solely from
markets where no particular source of PFAS contamination was identified (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-Katarzynska et al., 2019). Across studies, PFBS
detections were found in seafood; 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
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.

Two of the 12 European studies examined both market-bought and fresh-caught fish, and PFBS
was detected in seafood from both sources (Vassiliadou et al., 2015; Yamada et al., 2014).
Yamada et al. (2014) found higher PFBS in fresh-caught river fish samples (0.16 ng/g ww
maximum) versus fresh or frozen market samples (0.03 ng/g ww maximum) in France.
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.

In summary, in Europe and North America, PFBS has been detected in multiple food types,
including fruits, vegetables, meats, seafoods, and other fats. Several large-scale fish tissue
sampling efforts conducted by EPA and others indicate that fish consumption may be an
important PFBS exposure source. Future large-scale sampling efforts by FDA and others may
help to similarly elucidate PFBS concentrations in other food types. 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.

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

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route of exposure (ATSDR, 2021; U.S. EPA, 2021a). 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).

Five studies in Europe (conducted in Poland, Norway, Greece, Czech Republic, and Germany)
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 LOQ 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), and aluminum foil bags/wrappers (n = 14). The study concluded that the use of
perfluorinated compound alternatives such as fluorophosphates and fluorinated polyethers in the
local manufacturing process potentially explains the low levels of other PFAS (i.e.,
perfluorobutanoic acid [PFBA], perfluorohexanoic acid [PFHxA], perfluoroheptanoic acid
[PFHpA], perfluorononanoic acid [PFNA], perfluorodecanoic acid [PFDA], and
perfluorododecanoic acid [PFDoDA]) detected in the sampled FCMs. 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.

Overall, the few available studies conducted in the United States and Europe indicate 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.

3.3.1.2 Consumer Products

Consumer products could also be a source of PFBS exposure as noted in Section 1.3. 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.,

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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. Additionally, two European studies analyzed
commercially available AFFF products which have been formulated with PFAS and are
associated with elevated levels of these chemicals in environmental media (Favreau et al., 2016;
Fteisseter et al., 2019).

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 non-
woven medical garments, floor waxes, membranes for apparel, and thread-seal ant 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 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 (1 pair of outdoor trousers, 7 jackets, 4
fabrics for outdoor clothes, 1 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
(PFAA) levels of 5-fold or more. For 8 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 3 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 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 (j,g/m2), ski wax
samples (up to 3.1 micrograms per kilogram [[j,g/kg]), leather samples (up to 120 (J,g/kg), and
gloves (up to 2 (J,g/kg). Favreau et al. (2016) analyzed the occurrence of 41 PFAS in a wide

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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 |ig/kg found in a used 1999 screen
associated with WEEE.

PFBS was also evaluated in AFFFs in Switzerland (Favreau et al., 2016) and Norway (Fteisseter
et al., 2019). In currently commercially available AFFFs from Switzerland, PFBS was detected
in 11% of samples (n = 35) with a maximum concentration of 0.1 ppm (Favreau et al., 2016). In
AFFFs used at a firefighting training facility in Norway, PFAS concentrations in 1:100 diluted
AFFF were predominately PFOS (88.7%). PFBS contributed to 1.2% of the concentration of the
23 total PFAS tested in the diluted foam, with a concentration of 1,400,000 ng/L (Fteisseter et al.,
2019).

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.

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3.3.1.3 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.
Eighteen 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 B-5). 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 >
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 entry ways to the yard. The DFs for PFBS were 16% and
11% for living rooms and entry ways, 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.

3 J J.4 Air

PFAS have been released to air from WWTPs, waste incinerators, and landfills (U.S. EPA,
2016a). 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, EPA's Toxic
Release Inventory did not report release data for PFBS in 2020 (U.S. EPA, 2022a). In addition,
PFBS is not listed as a hazardous air pollutant (U.S. EPA, 2022b).

Indoor Air

Three studies in Europe, conducted in Norway (Barber et al., 2007), Spain (Jogsten et al., 2012),
and Ireland (Harrad et al., 2019), analyzed the occurrence of PFBS in indoor air samples.

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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; PFOS and PFBA were the
only detected PFCs in these indoor air samples.

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.

Ambient Air

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 PFAS
impact. 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 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 non-industrial 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 (PM2.5) in air
samples collected from five stations located in Alicante province, Spain (3 residential, 1 rural, 1
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 method detection limit.

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).

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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.

3.35 Soil

PFBS can be released into soil from manufacturing facilities, industrial uses, fire/crash training
sites, and biosolids containing PFBS (ATSDR, 2021, U.S. EPA, 2021a). 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 B-6). 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 (Nickerson et al., 2020). The study of the other fire training site measured PFBS pre-
treatment (0.61-0.6.4 ng/g) and post-treatment (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. Grannestad 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

30


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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). As mentioned, 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.

3.3.2 RSC Determination

In summary, based on the physical properties, detected levels, and available exposure
information for PFBS, multiple non-drinking water sources (seafood [including fish and
shellfish]) and other foods including vegetables, indoor air, and some consumer products) are
potentially significant exposure sources. Following the Exposure Decision Tree within EPA's
2000 Methodology for Deriving Ambient Water Quality Criteria for the Protection of Human
Health (U.S. EPA, 2000a), significant potential sources other than drinking water ingestion were
identified (Box 8A in the Decision Tree). However, information is not available to quantitatively
characterize the relative exposure contributions from the non-drinking water sources (Box 8B in
the Decision Tree, U.S. EPA, 2000a).

EPA also considered the exposure information specifically for the identified sensitive population.
The identified sensitive lifestage, based on the critical study and effect, is women of childbearing
age (13 to <50 years) who may be or become pregnant. However, the literature search did not
identify non-drinking water exposure information specific to women of childbearing age that
could be used quantitatively to derive an RSC. Since neither the available data for the general
population (all ages) nor the sensitive population enabled quantitative characterization of relative
exposure sources and routes, EPA relied on an RSC of 20% (see Section 2.2.4.2 above; U.S.
EPA, 2000a), which means that 20% of the exposure equal to the RfD is allocated to drinking
water and the remaining 80% is reserved for other potential exposure sources such as food,
indoor air, and some consumer products.

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4 J Lifetime Noncancer Health Advisory Derivation

The lifetime noncancer HA for PFBS is calculated as follows:

RfD

Lifetime HA

( KtL) \

= 	 * RSC

VDWI-BW/

Lifetime HA

/0.0003 , m,g

	^1*0.2

\ 0.0354, A
\	kg-day

Lifetime HA = 0.0017 —grounded to 0.002 )

_

L

ng

= 2,000 -2

J_j

(Eq. 3)

EPA is issuing a lifetime noncancer drinking water HA for PFBS of 2,000 ng/L (ppt). The
critical health effect on which the chronic RfD used to calculate the lifetime HA is based (i.e.,
decreased serum levels of the T/dn newborn mice) resulted from PFBS exposure during a
developmental life stage. In Feng et al. (2017), developmental effects occurred atPND 1 and
were sustained through pubertal (PND 30) and adult periods (PND 60). This is consistent with
the potential for long-term health consequences of gestational-only PFBS exposure and suggests
that gestation is at least one critical window for PFBS. EPA's risk assessment guidelines for
developmental toxicity indicate that adverse effects can result from even brief exposure during a
critical period of development (U.S. EPA, 1991). Therefore, the lifetime HA for PFBS of 2000
ng/L and the chronic RfD from which it is derived are considered applicable to short-term PFBS
exposure scenarios (including during pregnancy) as well as lifetime exposure scenarios via
drinking water. This lifetime HA applies to PFBS (CASRN 375-73-5), K+PFBS (CASRN 29420-
49-3), and PFBS- (CASRN 45187-15-3).

5 J Analytical Methods

EPA developed two liquid chromatography/tandem mass spectrometry (LC/MS/MS) analytical
methods to quantitatively monitor drinking water for targeted PFAS that include PFBS: EPA
Method 533 (U.S. EPA, 2019b) and EPA Method 537.1, Version 2.0 (U.S. EPA, 2020b). The
methods discussed below can be used to accurately and reasonably quantitate PFBS at ng/L
levels that are three orders of magnitude below the PFBS lifetime HA of 2000 ng/L.

EPA Method 533 monitors for 25 select PFAS with published measurement accuracy and
precision data for PFBS in reagent water, finished groundwater, and finished surface water and a
single laboratory-derived MRL or approximate quantitation limit for PFBS at 3.5 ng/L (0.0035
|ig/L), For further details about the procedures for this analytical method, please see Method

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533: Determination of Per- and Polyfluoroalkyl Substances in Drinking Water by Isotope
Dilution Anion Exchange Solid Phase Extraction and Liquid Chromatography/Tandem Mass
Spectrometry (U.S. EPA, 2019b).

EPA Method 537.1 (representing an update to EPA Method 537 [U.S. EPA, 2009b]) monitors
for 18 select PFAS with published measurement accuracy and precision data for PFBS in reagent
water, finished groundwater, and finished surface water and a single laboratory-derived MRL or
approximate quantitation limit for PFBS at 6.3 ng/L (0.0063 |ig/L). For further details about the
procedures for this analytical method, please see Method 537.1: Determination of Selected Per-
and Polyfluorinated Alky I Substances in Drinking Water by Solid Phase Extraction and Liquid
Chromatography/Tandem Mass Spectrometry (LCZMSZMS) (U.S. EPA, 2020b).

Drinking water analytical laboratories have different performance capabilities dependent upon
their instrumentation (manufacturer, age, usage, routine maintenance, operating configuration,
etc.) and analyst experience. Some laboratories will effectively generate accurate, precise,
quantifiable results at lower concentrations than others. Organizations leading efforts that include
the collection of data need to establish data quality objectives (DQOs) to meet the needs of their
program. These DQOs should consider establishing reasonable quantitation limits that
laboratories can routinely meet, without recurring quality control (QC) failures that will
necessitate repeating sample analyses, increase costs, and potentially reduce laboratory capacity.
Establishing a quantitation limit that is too high may result in important lower-concentration
results being overlooked.

EPA's approach to establishing DQOs within the UCMR program serves as an example. EPA
established MRLs for UCMR 5,15 and requires laboratories approved to analyze UCMR samples
to demonstrate that they can make quality measurements at or below the established MRLs. EPA
calculated the UCMR 5 MRLs using quantitation-limit data from multiple laboratories
participating in an MRL-setting study. The laboratories' quantitation limits represent their lowest
concentration for which future recovery is expected, with 99% confidence, to be between 50 and
150%. The UCMR 5-derived and promulgated MRL for PFBS is 0.003 |ig/L (3 ng/L).

6 J Treatment Technologies

This section summarizes available drinking water treatment technologies that have been
demonstrated to remove PFBS from drinking water, but it is not meant to provide specific
operational guidance or design criteria. High-pressure membrane processes such as nanofiltration
(NF) and reverse osmosis (RO) are generally effective at removing organic solutes and dissolved
ions and have been shown to successfully reduce or remove PFBS from drinking water
(Appleman et al., 2014). NF generally removes 20-10% of PFBS (Jin et al., 2021), although
93% (Appleman et al., 2013) and 99.8% (U.S. EPA, 2021e) removal have been reported with
NF. The amount of contaminant removed by membranes is referred to as a rejection rate; RO
tends to have a higher rejection rate than NF. Direct filtration NF and RO membranes have been
successful in removing PFBS at full-scale water treatment works to below the 3 ppt EPA UCMR
5 reporting limit (Appleman et al., 2014; Konradt et al., 2021; Liu et al., 2021; Quinones and
Snyder, 2009; Thompson et al., 2011). Absorption-based NF and RO membranes have had

15 Information about UCMR 5 is available at https://www.epa.gov/dwcira/fiflh-unregulated-contamiiiant-monitoriiig-rule

33


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success with PFBS treatment at laboratory scale (Zhang et al., 2019). Hybrid membrane
processes, such as applying direct-current electrical fields or photocatalysts across lower pressure
membranes, have had success with other short-chain sulfonates at laboratory scale (Tsai et al.,
2010; Urtiaga, 2021). For more information about hybrid membrane processes, see Soriano et al.
(2020) or (2017). Installing high-pressure membranes may have additional benefits on finished
water quality by removing other contaminants and disinfection byproduct precursors. Sorption-
based processes such as activated carbon and ion exchange have been shown to remove PFBS in
drinking water to below the EPA method reporting limit of 3 ppt for UCMR 5; however, the
media usage rate is higher than for other PFAS with longer carbon backbones (McCleaf et al.,
2017; Murray et al., 2021). Information about PFBS treatment efficacy with sorption-based
processes is still emerging; more information about the suitability of these technologies is
expected to be available in the future. Most other treatment processes are viewed as not
sufficiently effective or cost efficient to reduce PFBS concentrations in drinking water. For
example, coagulation, flocculation, sedimentation, and biologically active carbon filtration are
generally ineffective at removing PFBS (Quinones and Snyder, 2009; Sun et al. 2016).

Ozonation has increased concentrations of PFBS at full-scale water treatment plants (WTPs),
possibly due to PFAS precursor compound oxidation (Sun et al., 2016). Boiling water will
concentrate PFBS and should not be considered as an emergency action.

Non-treatment PFBS management practices such as changing source waters, source water
protection, or consolidation are also viable options for reducing PFBS concentrations in finished
drinking water. One resource for protecting source water from PFAS, including PFBS, is the
PFAS - Source water Protection Guide and Toolkit (ASDWA, 2020), which shares effective
strategies for addressing PFAS contamination risk in source waters. Source water protection is
particularly important since natural attenuation is not a valid PFBS management strategy. PFBS
will not degrade by abiotic reaction mechanisms such as hydrolysis and photolysis under
environmental conditions (Lassen et al., 2013; NICNAS, 2005). Likewise, Quinete et al. (2010)
studied biotic PFBS degradability using the manometric respirometry test (OECD, 1992b) and
the closed-bottle test (OECD, 1992a) with River Rhine water as inoculum; PFBS did not show
signs of biodegradation in either test.

NF and RO are high pressure processes where water is forced across a membrane. The water that
transverses the membrane is known as permeate or produce, and has few solutes left in it; the
remaining water is known as concentrate, brine, retentate, or reject water and forms a waste
stream with concentrated solutes. The main PFBS removal mechanisms in NF and RO are steric
exclusion, solution-diffusion, and electrostatic interaction (Jin et al., 2021). NF has a less dense
active layer than RO, which enables lower operating pressures but also makes it less effective at
removing contaminants. Higher operating pressures and initial flux generally enhance removal.
Temperature and pH are also significant parameters affecting performance. In general, organic
NF membranes have lower operating costs and easier processing than inorganic membranes
while maintaining appropriate robustness for PFBS treatment (Jin et al., 2021). NF and RO tend
to have high operating expenses, use significant amounts of energy, and generate concentrate
waste streams which require disposal. Generally, NF and RO require pre- and posttreatment
processes.

PFBS removal fluxes are generally around 40 liters per square meter per hour (L/[m2hr]) at
about 0.7 megapascal (MPa) operating pressure (Wang et al., 2018). Temperature can

34


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dramatically impact flux; it is common to normalize flux to a specific reference temperature for
operational purposes (U.S. EPA, 2005c). It is also common to normalize flux to pressure ratios to
identify productivity changes attributable to fouling (U.S. EPA, 2005c). It is important to note
that water may traverse the membranes from outside-in or inside-out; different system
configurations operating at the same flux produce differing quantities of finished water. This
means that membrane systems with differing configurations cannot be directly compared based
on flux. Total flow per module and cost per module are more important decision support
indicators for capital planning.

High-pressure membranes may have effects when added onto a well-functioning treatment train.
For instance, high-pressure membranes may remove beneficial minerals and increase corrosivity.
Increased water corrosivity may need to be addressed through corrosion treatment modifications
and water may require mineralization. For more information, see AWWA (2007).

6.1	Point-of-Use Devices for Individual Household PFBS Removal
Although the focus of this section is the different available options for removal of PFBS at
DWTPs, centralized treatment technologies can also be used in a decentralized fashion as point-
of-entry (POE) (where the distribution system meets a service connection) or point-of-use (POU)
(at a specific tap or application) treatment in cases where centralized treatment is impractical or
individual consumers wish to further reduce their individual household risks. Many home
drinking water treatment units are certified by independent third-party accreditation
organizations using American National Standards Institute (ANSI) standards to verify
contaminant removal claims. NSF International has developed a protocol for NSF/ANSI
Standard 58 (RO) that establishes minimum requirements for materials, design, construction, and
performance of POU systems (NSF/ANSI, 2021). Currently, these standards provide certification
procedures for PFOA and PFOS removal in drinking water to below EPA's 2016 PFOA and
PFOS HA level of 70 ppt. When properly maintained, these systems may reduce other PFAS,
including PFBS, although removal should not be automatically inferred for PFAS not specified
within the protocol. PFBS removal by faucet filters has reportedly averaged 94%, whereas
pitcher filters had an average of 65% removal, refrigerator filters 29%, single-stage under-sink
filters 84%), two-stage filters > 92%, and RO filters 94% (Herkert et al., 2020). PFBS specific
certification procedures may be developed in the future by voluntary consensus standards
organizations. Individuals interested in POU or POE treatment should check with the
manufacturers of these devices as to whether they have been independently certified for the
reduction of PFBS levels in drinking water.

6.2	Treatment Technologies Summary

Non-treatment PFBS management options, such as changing source waters, source water
protection, or consolidation, are viable strategies for reducing PFBS concentrations in finished
drinking water. Should treatment be necessary, NF along with RO are the best means for
removing PFBS from drinking water and can be used in central treatment plants or in POU/POE
applications. Sorption processes such as activated carbon or ion exchange may successfully
remove PFBS, but with lower efficacy than PFAS with a longer carbon backbone such as PFOS.
PFBS treatment technologies often require pre- as well as post-treatment and may help remove
other unwanted contaminants and disinfection byproduct precursors. These treatment processes
are separation technologies and produce waste streams with PFBS on or in them.

35


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7 J Consideration of Noncancer Health Risks from PFAS Mixtures

EPA recently released a Draft Framework for Estimating Noncancer Health Risks Associated
with Mixtures of Per- and Polyfluoroalkyl Substances (PFAS) (U.S. EPA, 2021f) that is currently
undergoing Science Advisory Board (SAB) review. That draft document describes a flexible,
data-driven framework that facilitates practical component-based mixtures evaluation of two or
more PFAS based on current, available EPA chemical mixtures approaches and methods (U.S.
EPA, 2000b). Examples are presented for three approaches—Hazard Index (HI), Relative
Potency Factor (RPF), and Mixture BMD—to demonstrate application to PFAS mixtures. To use
these approaches, specific input values and information for each PFAS are needed or can be
developed. These approaches may help to inform PFAS evaluation(s) by federal, state, and tribal
partners, as well as public health experts, drinking water utility personnel, and other stakeholders
interested in assessing the potential noncancer human health hazards and risks associated with
PFAS mixtures.

The HI approach, for example, could be used to assess the potential noncancer risk of a mixture
of four component PFAS for which HAs, either final or interim (iHA), are available from EPA
(PFOA, PFOS, GenX chemicals [hexafluoropropylene oxide dimer acid and its ammonium salt],
and PFBS). In the HI approach described in the draft framework (U.S. EPA 202If), 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) (U.S. EPA, 1986).
The HI is dimensionless, so in the HI formula, E and the RfV must be in the same units (Eq. 6).
In the context of PFAS in drinking water, a mixture PFAS HI can be calculated when health-
based water concentrations (e.g., HAs, Maximum Contaminant Level Goals [MCLGs]) for a set
of PFAS are available or can be calculated. In this example, HQs are calculated by dividing the
measured component PFAS concentration in water (e.g., expressed as ng/L) by the relevant HA
(e.g., expressed as ng/L) (Eqs. 7, 8). The component chemical HQs are then summed across the
PFAS mixture to yield the mixture PFAS His based on interim and final HAs.

»'=2>=S

i=l	i=l

(Eq. 6)

HI = HQpfqa + HQpFos + HQGenX + HQPFBS

(Eq. 7)

/[PFBSwater]\

V [pfbsha] J

(Eq. 8)
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

/[PFOAwater]\ /[PFOSwater]\ /[GenXwater]
" V [PFOAiHA] J [ [PFOSiHA] J [ [GenXHA]

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RfV = human health-based toxicity value for component (i) PFAS
HQpfas= hazard quotient for a given PFAS
[PFASwater] = concentration of a given PFAS in water
[PFASha] = HA value, interim or final, for a given PFAS

In cases when the mixture PFAS HI is greater than 1, this indicates an exceedance of the health
protective level and indicates potential human health risk for noncancer effects from the PFAS
mixture in water. When component health-based water concentrations (in this case, HAs) are
below the analytical method detection limit, as is the case for PFOA and PFOS, such individual
component HQs exceed 1, meaning that any detectable level of those component PFAS will
result in an HI greater than 1 for the whole mixture. Further analysis could provide a refined
assessment of the potential for health effects associated with the individual PFAS and their
contributions to the potential joint toxicity associated with the mixture. For more details of the
approach and illustrative examples of the RPF approach and Mixture BMD approaches please
see U.S. EPA (202If).

8 J Health Advisory Characterization

EPA is issuing a lifetime noncancer drinking water HA for PFBS of 2,000 ng/L or 2,000 ppt
based on the best available science. This is the first HA for PFBS. The PFBS HA is considered
applicable to both short-term and chronic risk assessment scenarios because the critical effect
identified for PFBS can result from developmental exposure and leads to long-term adverse
health effects (Feng et al., 2017). The input values for the HA include 1) the chronic RfD which
was developed in the toxicity assessment for PFBS (U.S. EPA 2021a); 2) the RSC based on
exposure information collected from a literature search and following EPA's Exposure Decision
Tree (U.S. EPA, 2000a) and presented herein; and 3) the DWI-BW, described herein, selected
for the sensitive population or lifestage. The PFBS toxicity assessment was published after
rigorous scientific review, including internal and external review, and public comment.

Some of the uncertainties associated with the PFBS noncancer lifetime HA are due to data gaps.
The PFBS toxicity assessment, which was the basis for the chronic RfD used to derive the HA,
performed a systematic literature search and identified a limited number of studies examining
health effects after PFBS exposure (U.S. EPA, 2021a). The toxicity assessment literature search
did not identify available chronic studies or cancer studies for PFBS. Only a small number of
human studies per health outcome category were identified. The identified animal studies of
repeated-dose PFBS exposure used K+PFBS as the tested substance and only examined
noncancer effects. Further, since neurodevelopmental effects are of particular concern when
perturbations in thyroid hormone occur during development, studies evaluating
neurodevelopmental effects following PFBS exposure during development are needed (U.S.
EPA, 2021a). Mechanistic studies were assessed as part of the systematic literature review but
mechanism(s) of toxicity for PFBS for the various health outcomes have not been established.

Based on the data gaps and limitations described above, there is some uncertainty about whether
the most sensitive population or life stage for PFBS exposure has been identified. Results of the
literature search for information that could inform RSC determination for PFBS indicate that
there is significant exposure from media other than drinking water, but the available data do not
allow for quantitative characterization of the contributions of non-drinking water exposures. This

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final HA is based on a recent toxicity assessment and recent literature searches of the publicly
available scientific information regarding health effects, exposure, analytical methods, and
treatment technologies for PFBS.

8.1 Comparative Analysis of Exposure Factors for Different Populations
The exposure duration in the critical study identified for PFBS in the toxicity assessment (U.S.
EPA, 2021a) is throughout gestation which suggests that pregnant women and their developing
embryo and fetus represent a sensitive life stage. In addition to drinking water exposure to
pregnant women (and their developing embryo and fetus), the gestational exposure window is
relevant to drinking water exposure to women of childbearing age (13 to < 50 years) who may be
or become pregnant (Table 5).

EPA compared the impact of using the DWI-BW for the 90th percentile for the general
population (all ages) with the DWI-BWs for the potentially sensitive populations identified,
women of childbearing age and pregnant women on the HA value (Table 6). All three HA values
are the same when rounded to one significant figure (i.e., all are 0.002 ppm). This indicates that
the lifetime noncancer HA developed for PFBS based on the selected DWI-BW for women of
childbearing age is protective of the 90th percentile of all ages of the general population.

Table 6. Comparison of HA Values Using EPA Exposure Factors for Drinking Water
Intake for Different Candidate Populations

Population

DWI-BW

(L/kg bw-day)

HA

calculated/HA
rounded to one
significant figure

Description of Exposure
Metric

Source

Pregnant women

0.0333

0.00180/
0.002 ppm

90th percentile direct and
indirect consumption of
community water,
consumer-only two-day
average.

2019 Exposure
Factors Handbook
Chapter 3, Table 3-
63, NHANES 2005-
2010 (U.S. EPA,
2019a)

Women of
childbearing age

0.0354

0.00169/
0.002 ppm

90th percentile direct and
indirect consumption of
community water,
consumer-only two-day
average, 13 to < 50 years.

General

population, all ages

0.0338

0.00177/
0.002 ppm

90th percentile direct and
indirect consumption of
community water,
consumer-only two-day
average, all ages.

2019 Exposure
Factors Handbook
Chapter 3, Table 3-
21, NHANES 2005-
2010 (U.S. EPA,
2019a)

Notes'. L/kg bw-day = liters of water consumed per kilogram body weight per day. The DWI-BW used to calculate the PFBS
lifetime HA is in bold.

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Bach, C., X. Dauchy, V. Boiteux, A. Colin, J. Hemard, V. Sagres, C. Rosin, and J.F. Munoz.
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Zafeiraki, E., D. Costopoulou, I. Vassiliadou, L. Leondiadis, E. Dassenakis, R. Hoogenboom,
and S.P.J, van Leeuwen. 2016a. Perfluoroalkylated substances (PFASs) in home and
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Zafeiraki, E., I. Vassiliadou, D. Costopoulou, L. Leondiadis, H.A. Schafft, R. Hoogenboom, and
S.P.J, van Leeuwen. 2016b. Perfluoroalkylated substances in edible livers of farm
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Chemosphere 156:280-285. doi: 10.1016/j.chemosphere.2016.05.003.

Zhang, X., R. Lohmann, C. Dassuncao, X.C. Hu, A.K. Weber, C.D. Vecitis, and E.M.

Sunderland. 2016. Source attribution of poly- and perfluoroalkyl substances (PFASs) in
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Zhang, D.Q., W.L. Zhang, and Y.N. Liang. 2019. Adsorption of perfluoroalkyl and

polyfluoroalkyl substances (PFASs) from aqueous solution- a review. Science of the
Total Environment 694:133606.

Zhao, Z., Z. Xie, J. Tang, R. Sturm, Y. Chen, G. Zhang, and R. Ebinghaus. 2015. Seasonal
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Zheng, G., B.E. Boor, E. Schreder, and A. Salamova. 2020. Indoor exposure to per- and
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Appendix A: Relative Source Contribution - Literature Search and
Screening Methodology

Information on all exposure sources and routes for perfluorobutane sulfonic acid (PFBS )was
gathered through a literature search in a manner consistent with the Office of Science and
Technology's (OST's) process the collection of information for relative source contribution
(RSC) derivation. In this process, a literature search of both the peer reviewed and gray literature
for the chemical of interest was conducted. All of the primary studies that were identified from
the search are final documents or articles.

In 2020, U.S. Environmental Protection Agency's (EPA's) Office of Research and Development
(ORD) conducted a broad literature search to evaluate evidence for pathways of human exposure
to eight per- and polyfluoroalkyl substances (PFAS): perfluorooctanoic acid (PFOA),
perfluorooctanesulfonic acid (PFOS), perfluorobutanoic acid (PFBA), perfluorobutane sulfonic
acid (PFBS), perfluorodecanoic acid (PFDA), perfluorohexanoic acid (PFHxA), perfluorohexane
sulfonic acid (PFHxS), and perfluorononanoic acid (PFNA). This search was not date limited
and spanned the information collected across the Web of Science, PubMed, and ToxNet/ToxLine
(now ProQuest) databases. The results of the PFBS literature search of publicly available sources
are available through EPA's Health & Environmental Resource Online website at
https://hero.epa.eov/hero/index.cfm/proiect/paee/proiect id/2610.

The 654 literature search results for PFBS were imported into SWIFT-Review (Sciome, LLC,
Research Triangle Park, NC) and filtered through the Evidence Stream tags to identify human
studies and non-human (i.e., those not identified as human) studies. Human studies were further
categorized into seven major PFAS pathway categories (Cleaning Products, Clothing,
Environmental Media, Food Packaging, Home Products/Articles/Materials, Personal Care
Products, and Specialty Products) plus an additional category for Human Exposure Measures.
Non-human studies were grouped into the same seven major PFAS pathway categories, except
that the Environmental Media category did not include soil, wastewater, or landfill. Only studies
published between 2003 and 2020 were considered. Application of the SWIFT-Review tags
identified 343 peer-reviewed papers matching these criteria for PFBS.

After this 2020 literature search was conducted, the 343 articles were screened to identify studies
reporting measured occurrence of PFBS in human matrices and media commonly related to
human exposure (human blood/serum/urine, drinking water, food, food contact materials,
consumer products, indoor dust, indoor and ambient air, and soil). For this synthesis, additional
screening was conducted to identify studies relevant to surface water (freshwater only) and
groundwater using a keyword16 search for water terms.

Following the Populations, Exposures, Comparators, and Outcomes (PECO) inclusion criteria
outlined in Table A-l, the title and abstract of each study were independently screened for
relevance by two screeners using litstream™. A study was included as relevant if it was unclear
from the title and abstract whether it met the inclusion criteria. When two screeners did not agree
if a study should be included or excluded, a third reviewer made a final decision. The title and
abstract screening of and of this synthesis resulted in 191 unique studies being tagged as relevant

16 Keyword list: water, aquifer, direct water, freshwater, fresh water, groundwater, groundwater, indirect water, lake, meltwater,
melt water, natural water, overland flow, recreation water, recreational water, river, riverine water, riverwater, river water,
springwater, spring water, stream, surface water, total water, water supply

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(i.e., having data on occurrence of PFBS in exposure media of interest) that were further
screened with full-text review using the same inclusion criteria. After additional review of the
evidence collected by ORD, 87 studies originally identified for other PFAS also contained
information relevant to PFBS. Based on full-text review, 147 studies were identified as having
relevant, extractable data for PFBS from the United States, Canada, or Europe for environmental
media, not including studies with only human biomonitoring data. Of these 147 studies, 130
were identified from the ORD literature search, where primary data were extracted into a
comprehensive evidence database. Parameters of interest included sampling dates and locations,
numbers of collection sites and participants, analytical methods, limits of detection and detection
frequencies, and occurrence statistics. Seventeen of the 147 studies were identified in this
synthesis as containing primary data on only surface water and/or groundwater.

Table A-l. Populations, Exposures, Comparators, and Outcomes (PECO) Criteria

PECO Element

Inclusion Criteria

Population

Adults and/or children in the general and impacted populations from the
United States, Canada, or Europe

Exposure

Primary data from peer-reviewed studies collected in any of the following
media: ambient air, consumer products, drinking water, dust, food, food
packaging, groundwater3, human blood/serum/urine, indoor air, landfill,
sediment, soil, surface water3 (freshwater), wastewater/biosolids/sludge

Comparator

Not applicable

Outcome

Measured concentrations of PFBS (or measured emissions from food
packaging and consumer products only)

Note:

a Surface water and groundwater were not included as relevant media in ORD's literature search. Studies were re-screened for
these two media in this synthesis.

The evidence database additionally identified 18 studies for which the main article was not
available for review. As part of this synthesis, 17 of the 18 studies could be retrieved. An
additional three references were identified through gray literature sources that were included to
supplement the search results. The combined 20 studies underwent full-text screening using the
inclusion criteria in Table A-l. Based on full-text review, four studies were identified as
relevant.

Using the screening results from the evidence database and this synthesis, a total of 151 studies
were identified as relevant and are summarized below.

To supplement the primary literature database, EPA also searched the following gray literature
sources for information related to relative exposure of PFBS for all potentially relevant routes of
exposure (oral, inhalation, dermal) and exposure pathways relevant to humans:

•	U.S. EPA. 2021a. Human Health Toxicity Values for Perjluorobutane Sulfonic Acid
(CASRN 375-73-5) and Related Compound Potassium Perfluorobutane Sulfonate
(CASRN 29420-49-3).

•	AT SDR's Toxicological Profiles

•	Centers for Disease Control's national reports on human exposures to environmental
chemicals

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•	EPA's CompTox Chemicals Dashboard

•	EPA's fish tissue studies

•	EPA's Toxics Release Inventory

•	EPA's Unregulated Contaminant Monitoring Rule (UCMR) data

•	Relevant documents submitted under Toxic Substances Control Act (TSCA) and relevant
reports from U.S. EPA's Office of Chemical Safety and Pollution Prevention

•	FDA's Total Diet Studies and other similar publications from FDA, U.S. Department of
Agriculture, and Health Canada

•	National Oceanic and Atmospheric Administration's National Centers for Coastal Ocean
Science data collections

•	National Science Foundation direct and indirect food and/or certified drinking water
additives

•	PubChem compound summaries

•	Relevant sources identified in the RSC discussions (section 5) of EPA's Proposed
Approaches to the Derivation of a Draft Maximum Contaminant Level Goal for
Perfluorooctanoic Acid (PFOA)/Perfluorooctane Sulfonic Acid (PFOS) in Drinking
Water

•	Additional sources, as needed

EPA has included available information from these gray literature sources for PFBS relevant to
its uses, chemical and physical properties, and for occurrence in drinking water (directly or
indirectly in beverages like coffee, tea, commercial beverages, or soup), ambient air, foods
(including fish and shellfish), incidental soil/dust ingestion, and consumer products. EPA has
also included available information specific to PFBS on any regulations that may restrict PFBS
levels in media (e.g., water quality standards, air quality standards, food tolerance levels).

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Appendix R: Compilation of Data on PFRS Occurrence In
Environmental Media Collected from Primary Literature

This appendix includes tables resulting from the efforts to identify and screen primary literature
(i.e., peer-reviewed journal articles), described in Appendix A, as well as extract data that may
be relevant to informing the RSC derivation for PFBS.

Table B-l. Compilation of Studies Describing PFBS Occurrence in Drinking Water

Study

Location

Site Details

Results

North America

Bradley et al.
(2020)

United States
(Chicago, Illinois;
East Chicago,
Indiana)

Residential tap water (45
sites); treated, pre-
distribution tap water from
water filtration plants (4
sites)

Residential tap water: DFa 47%,
range = ND-0.8 ng/L

Pre-distribution tap water = DFa
75%, range = ND-0.5 ng/L

Hu et al. (2019)

United States
(national)

Archived tap water samples
(collected 1989-1990) from
225 homes of Nurses'
Health Study participants
(across 22 states)

DF 5%, median (range) = 0.20
(ND-2.97) ng/L

Boone et al. (2019)

United States
(national)

Treated water from 25
DWTPs; some locations
reportedly had known or
suspected sources of
wastewater in the source
water, but the study did not
identify which

DF 96%, median (range) = 1.17
(ND-11.9) ng/L

Dasu et al. (2017)

United States
(Ohio, Kentucky)

Tap water collected in
2003-2006 from 25 homes
of Health Outcomes and
Measures of the
Environment study
participants

DF 16%, range = ND-11.7 ng/L

Subedi et al. (2015)

United States (New
York)

Tap water (from outdoor
taps; 27 samples) from 4
homes around Skaneateles
Lake that use an enhanced
treatment unit for onsite
wastewater treatment

DF 7%, mean (range) = 0.44
(ND-0.48) ng/L

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Study

Location

Site Details

Results

Appleman et al.
(2014)

United States
(Wisconsin,
Oklahoma, Alaska,
California,
Alabama,
Colorado, Ohio,
Nevada,
Minnesota, New
Jersey)

Finished water from DWTPs
where source waters were
impacted by upstream
wastewater effluent
discharge

DF 100% (n=19), mean3 (range)
= 4.27 (0.43 - 37) ng/L

Scheretal. (2018)

United States
(Twin Cities
metropolitan
region, Minnesota)

Tap water from exterior taps
of homes near former 3M
PFAS production facility; 20
homes within and 3 homes
outside of the GCA (GCA
defined by well monitoring
conducted by Minnesota
Department of Health and
the Minnesota Pollution
Control Agency)

Within GCA: DF 0%
Outside GCA: DF 0%

Boone et al. (2014)

United States (New
Orleans, Louisiana)

Tap water from one home
when the river source water
was at a low stage (2.95 ft)
or a high stage (8.32 ft);
well water samples from
wells on a firefighting
training site that used AFFF
(3 wells sampled before
carbon adsorption treatment
and 1 well sampled after;
number of samples collected
per well not reported)

Tap water (low river stage): DF
100%, mean of primary and
duplicate = 14.15 ng/L

Tap water (high river stage): DF
100%, mean of 4 replicates =
2.12 ng/L

Well 1: DF NR, mean = 11.9
ng/L

Well 1 (after carbon adsorption
treatment): DF NR, mean =
9.09 ng/L

Well 2: DF NR, mean = 9.265
ng/L

Well 3: DF NR, mean = 29
ng/L

Lindstrom et al.
(2011)

United States
(Alabama)

Samples from 6 wells used
for drinking water located in
areas with historical land
application of
fluorochemical industry-
impacted biosolids

DFa 66%, mean (range) = 19.7
(ND-56.5) ng/L

Chow et al. (2021)

United States
(Baltimore,
Maryland
metropolitan area)

101 different non-
carbonated bottled water
products representing 66
brands, purchased from 19
different retail food and
beverage chains

DF 17%, median (range) = 0.25
(ND-1.44) ng/L

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Study

Location

Site Details

Results

Europe

Harrad etal. (2019)

Ireland (Dublin,
Galway, and
Limerick counties)

Bottled water (31) from
Galway city shops; tap water
(private supply) from 25
homes with private water
supplies; tap water (main
public supply) from 34
homes and 32 offices
(combined)

Bottled water: DF 29%, mean
(range) = 3.7 (ND-51) ng/L

Tap water (private supply): DF
0%

Tap water (main public supply):
DF 8%, mean (range) = 0.52
(ND-15.06) ng/L

Unlii Endirlik et al.
(2019)

Turkey (33
provinces)

Bottled water (26 samples
representing 18 different
brands, both plastic- and
glass-bottled); municipal tap
water (94 samples)

Bottled water: DF 8%, mean
(range) = 0.20 (ND-0.21) ng/L

Tap water: DF 87%, mean
(range) = 0.29 (ND-0.85) ng/L

Ciofi etal. (2018)

Italy (Tuscany)

8 drinking water samples
from various rural, urban,
and industrial districts of
Tuscany (origins not further
described, but latitudinal and
longitudinal coordinates for
sampling locations were
provided)

DF 0%

Le Coadou et al.
(2017)

France (national)

Bottled water (25 samples of
natural mineral water and 15
samples of spring water)

DF 2.5% (only one detection);
single detection value (range) =
1.4 (ND-1.4) ng/L

Shafique et al.
(2017)

Germany (Leipzig)

Tap water (2 samples) from
one location (authors'
research institute)

DF NR, mean =1.3 ng/L

Filipovic and
Berger(2015)

Sweden
(Bollebygd,
Bromma, Umea)

Tap water from four
WWTPs (4 or 5 samples
from each)

Bollebygd: DF 75%, mean =
0.015 ng/L

Norrvatten, Bromma: DF 100%,
mean = 1.33 ng/L

Stockholm Vatten, Bromma:
DF 100%, mean = 1.55 ng/L

Umea: DF 100%, mean = 0.035
ng/L

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Study

Location

Site Details

Results

Zafeiraki et al.
(2015)

Greece, the
Netherlands

Bottled water (5 samples
each from Greece and the
Netherlands); tap water
samples (37 samples from
the Netherlands and 43
samples from Greece)

Tap water:

Greece: DF 2.3% (only one
detection); single detection
value (range) = 0.7 (ND-0.7)
ng/L

The Netherlands: DF 35%,
median (range) = 7.6 (ND-
13.7) ng/L

Bottled water:

Greece: DF 0%

The Netherlands: DF 0%

Eschauzier et al.
(2013)

The Netherlands
(Amsterdam)

Hot water and tap water
from two different locations
(A and B), where A and B
originated from different
DWTPs; additional tap
water samples (n=4) from
cafes, universities, and
supermarkets

Hot water A: point = 3.3 ng/L

Tap water A: point = 3.2 ng/L

Hot water B: point =19 ng/L

Tap water B = 16 ng/L

Tap water (n=4): DF NR, mean
(range) = 16 (14-17) ng/L

Gellrich et al.
(2013)

Germany (Hesse,
Saxony Anhalt);
Switzerland; Czech
Republic

Bottled water; spring water;
tap water from homes

Bottled mineral water: DF 16%,
median (range) = 2.6 (ND-
13.3) ng/L

Spring water: DF 6%, median
(range) = 3.2 (ND-3.2) ng/L

Tap water: DF 42%, median
(range) = 2.7 (ND-5.8) ng/L

Eriksson et al.
(2013)

Denmark (Faroe
Islands)

Treated water from DWTPs
(source water from
Havnardal Lake or Kornvatn
Lake)

Havnardal Lake: DF 0%
Kornvatn Lake: DF 0%

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Study

Location

Site Details

Results

Boiteux et al.
(2012)

France (national)

Treated water from DWTPs
across two sampling
campaigns (41 samples in
first campaign, 69 samples
in second campaign)

First campaign (treated water
originating from surface water):

DF 46%, median = < 1 ng/L,
maximum = 3 ng/L

First campaign (treated water
originating from groundwater):

DF 40%, median = < 1 ng/L,
maximum = 3 ng/L

Second campaign (treated water
originating from surface water):

DF NR, range = ND-< 10 ng/L

Second campaign (treated water
originating from groundwater):

DF NR, range = ND-13 ng/L

Eschauzier et al.
(2012)

The Netherlands
(Amsterdam)

Finished water from DWTP
(n=5); tap water from 1
home

Finished water from DWTP:
DF NR, mean (range) = 20 (17-
24) ng/L

Tap water: point =19 ng/L

Llorca et al. (2012)

Germany, Spain

Mineral bottled water (2
samples from Germany, 4
samples from Spain); tap
water (84 samples from
Spain, 5 samples from
Germany); well water (2
samples from Spain, 0
samples from Germany)

Bottled water (both Germany
and Spain): DF 0%

Tap water:

Germany: DF 0%

Spain: DF 35%, mean (range) =
8.3 (ND-36 ng/L)

Well water (Spain): DF 0%

Ullahetal. (2011)

Belgium

(Antwerp);

Germany

(Schmallenberg);

Italy (Ispra); the

Netherlands

(Amsterdam);

Norway (Tromso):

Sweden

(Stockholm)

Tap water from seven
research institutes in six
European countries

Belgium: point = 2.94 ng/L

Germany: point = 0.092 ng/L

Italy: point = 0.502 ng/L

The Netherlands: DFa 100%,
mean3 (range) = 13.2 (7.61—
18.8) ng/L

Sweden: point = 0.955 ng/L
Norway: point = ND

Holzer et al. (2011)

Germany
(Sauerland)

Tap water (56 samples)
treated from Lake Mohne,
which became contaminated
by perfluorocompounds
through application of
polluted soil conditioner to
agricultural fields

DF 43%, mean (range) = 11
(ND-36) ng/L

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Study

Location

Site Details

Results

Ericson et al.
(2009)

Spain (5 regions of
Catalonia)

Tap water from 40 locations
identified as important
supply areas

Overall: DF 73%, mean (range)
= 4.52 (ND-69.43) ng/L

Barcelona: DF 86%, mean
(range) = 11.99 (ND-69.43)
ng/L

Girona: DF 57%, mean (range)
= 1.13 (ND-4.91) ng/L

Lleida: DF 43%, mean (range)
= 0.07 (ND-0.16) ng/L

Tarragona: DF 86%, mean
(range) = 0.32 (ND-0.55) ng/L

Terres de l'Ebre: DF 80%,
mean (range) = 0.45 (ND-1.28)
ng/L

Ericson et al.
(2008b)

Spain (Tarragona
Province)

Bottled water; municipal tap
water from public fountains
of most populated towns in
the province

Bottled water: DF 0%
Tap water: DF 0%

Pitteretal. (2020)

Italy (Veneto
region)

Treated water from DWTP
where its source water was
contaminated by PFAS
manufacturing plant

DF 89.5%, median (range) =
91.5 (ND-765.0) ng/L

Brandsma et al.
(2019)

The Netherlands
(Dordrecht)

Tap water from homes
within 50 km of
fluorochemical
manufacturing plant

DFa 100%, range = 2.5-11 ng/L

Li et al. (2018)

Sweden (Ronneby)

Finished water from
Brantafors DWTP, near
AFFF-contaminated military
airfield; finished water from
Karragarden DWTP

Brantafors: point =130 ng/L
Karragarden: DF 0%

Boiteux et al.
(2017)

France (northern)

Treated water from DWTPs
located 15-39 km
downstream of industrial
WWTP that processes raw
sewage from fluorochemical
manufacturing facility

DF 0%

Bach et al. (2017)

France (southern
region)

Treated water from two
DWTPs downstream of a
fluoropolymer
manufacturing facility

DF 0%

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Study

Location

Site Details

Results

Gebbink et al.
(2017)

The Netherlands

(Zwijndrecht,

Dordrecht,

Papendrecht,

Sliedrecht, Utrecht,

Wageningen)

Drinking water collected
from city halls in
municipalities close to
PFAS production plant
(D1-D4), at residential
home in Utrecht (D5), and at
the RIKILT institute in
Wageningen (D6)

D1: point = 3.4 ng/L
D2: point = 3.4 ng/L
D3: point =19 ng/L
D4: point = 2.3 ng/L
D5: point =1.0 ng/L
D6: point = 0.54 ng/L

Gyllenhammar et
al. (2015)

Sweden (Uppsala)

Finished water from
DWTPs; private well
(Klastorp) downstream of a
military airport using AFFF

Backlosa: DFa 9%, range =
ND-11 ng/L

Granby: DF 0%

Private well: DF 0%

Dauchy et al.
(2012)

France
(unspecified)

Treated water from DWTPs
located 15 km downstream
of fluorochemical
manufacturing facility

DF 0%

Weiss et al. (2012)

Germany
(Cologne)

Private well water 950 m
(Well A) and 2,000 m (Well
B) downstream of a fire
training area; Well A is
inside the contamination
plume.

Well A: DF 100%, mean3
(range) = 50 (20-100) ng/L

Well B: DFa 86%, range = ND-
20 ng/L

Multiple Continents

Kabore et al.
(2018)

Canada (Great
Lakes, St.
Lawrence River)

Tap water from homes (8
sites)

DF 100%, mean (range) = 0.5
(0.3-0.8) ng/L



Canada (rest of
Canada)

Tap water from homes (11
sites); bottled water (11
brands)

Tap water: DF 73%, mean
(range) = 0.1 (ND-0.5) ng/L

Bottled water: DFa 9%, range =
ND-0.23 ng/L



United States

(Illinois,

California)

Tap water from homes (2
sites)

DFa 50%; ND and 0.28 ng/L



Norway (Oslo)

Tap water from a home (1
site)

Point = 0.72 ng/L



France (Le Mans,
Paris, Guadeloupe
in French West
Indies)

Tap water from homes (3
sites)

DFa 67%, range = ND-0.32
ng/L

Notes: AFFF = aqueous film-forming foam; DF = detection frequency; DWTP = drinking water treatment plant; ft = feet; GCA =
groundwater contamination area; km = kilometer; m = meter; ND = not detected; ng/L = nanogram per liter; NR = not reported;
PFAS = per- and polyfluoroalkyl substances; WWTP = wastewater treatment plant; (ig/L = microgram per liter.
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 B-2. Compilation of Studies Describing PFBS Occurrence in Groundwater

Study

Location

Site Details

Results

North America

Lee et al. (2015)

United States
(California)

Samples from 5 urban
shallow groundwater wells
with wastewater
contamination

DFa 20%, range = ND-36.3
ng/L

Appleman et al.
(2014)

United States (New
Jersey)

Samples from 5 New Jersey
groundwater source waters
for PWSs impacted by
upstream wastewater
effluent discharge

DFa 100%, mean3 (range) =
2.4 (0.43-3.7) ng/L

Post et al. (2013)

United States (New
Jersey)

Raw water from 18 public
drinking water system
groundwater intakes

DF 6%, range = ND-6 ng/L

Steele et al.
(2018)

United States
(Alaska)

Military base contaminated
with PFAS from AFFF use
(4 wells sampled once per
month for 8 months)

DFa NR, range = ND-48 ng/L

Eberle et al.
(2017)

United States (Joint
Base Langley-
Eustis, VA)

Former fire training site, site
characterization and
pretreatment groundwater
samples

Site characterization: DF
100%, mean3 (range) = 3,700
(1,100-13,000) ng/L (10
wells)

Pretreatment: DF 100%, mean3
(range) = 3,400 (1,200-5,000)
ng/L (5 wells, 2 laboratory
samples/well)

Anderson et al.
(2016)

United States
(national)

Ten active U.S. Air Force
installations with historic
AFFF release

DF 78.26%, median of detects
(range) = 200 (ND-110,000)
ng/L

Moody et al.
(2003)

United States
(Oscoda, MI)

Groundwater plume at
former Wurtsmith Air Force
Base; firefighting training
area active from 1952 to
1993

DF 0%

Procopio et al.
(2017)

United States (New
Jersey)

Samples collected from
temporary wells in a small
area of an

industrial/business park
located within the
Metedeconk River
Watershed

DF 0%

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Study

Location

Site Details

Results

Lindstrom et al.
(2011)

United States
(Alabama)

Samples from 13 wells used
for purposes aside from
drinking water (e.g.,
livestock, watering gardens,
washing), located in areas
with historical land
application of
fluorochemical industry-
impacted biosolids

DFa 23%, mean (range) = 10.3
(ND-76.6) ng/L

Europe

Barreca et al.
(2020)

Italy (Lombardia
region)

Groundwater sampling
stations representative of
region

DF 18%a, concentrations NR

Boiteux et al.
(2012)

France (national)

Raw water from 2 sampling
campaigns of DWTPs, some
sites possibly affected by
industrial or commercial
releases

DF 4%, range = ND-9 ng/L

Loos et al. (2010)

23 European
countries

Monitoring stations were not
necessarily representative of
surrounding area or
contaminated

DF 15.2%, range = ND-25
ng/L

Gobelius et al.
(2018)

Sweden (national)

Sampling locations selected
based on potential vicinity
of PFAS hot spots and
importance as a drinking
water source area

DF 26%a (triplicate samples
removed), range = ND-22
ng/L

Dauchy et al.
(2012)

France (unspecified)

Raw water from 2 DWTPs
supplied by alluvial wells;
DWTPs located 15 km
downstream of
fluorochemical
manufacturing facility

DFa 40%, range = ND-4 ng/L

Hoisacter et al.
(2019)

Norway
(unspecified)

Samples from 19 sampling
campaigns of 5 pumping
wells placed to intercept a
groundwater contamination
plume originating from a
firefighting training facility
that ceased usage of PFAS-
and fluorotelomer-based
AFFF 15 years prior

Detections reported but DF
and concentrations not
provided

Dauchy et al.
(2019)

France (unspecified)

Samples collected over 2
campaigns from 6 areas (13
monitoring wells) of a
firefighter training site

DFa 77%, range = ND-750
ng/L

78


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Study

Location

Site Details

Results

Dauchy et al.
(2017)

France (unspecified)

Samples collected near 3
sites (A, C, D) impacted by
the use of AFFF. Site A
results describe 1 sampling
location with 2 sampling
events. Site C results
describe a single sampling
location and event. Site D
results describe 5 sampling
locations, each with a single
sampling event

Site A: DFa 100% mean3 = 8
ng/L

Site C: point = 6 ng/L

Site D: DFa 20%, range = ND-
59 ng/L

Gyllenhammar et
al. (2015)

Sweden (Uppsala)

Samples from local aquifers
extracted by 21 production
wells, 6 observation wells or
1 private well located in the
vicinity of a potential AFFF
point source (military
airport). Results for all well
sites were not provided.

Site 1 (production well): DF
0% (n = NR)

Site 3 (observation wells): DF
100%, median =100 ng/L (n =

3)

Site 5 (observation well): DF
0% (n = NR)

Site 6 (production well): DF
0% (n = NR)

Site 7 (observation well): DF
100%, median = 35 ng/L (n =

3)

Site 8 (production well): DFa
91%, median =13 ng/L (n =
103)

Site 10 (production well): DFa
2%, median = ND (n = 50)







Wagner et al.
(2013)

Germany
(unspecified)

Samples (n = 3) taken
downstream from a site
contaminated by AFFF from
firefighting activities

DFa 100%, concentrations NR

Notes: AFFF = aqueous film-forming foam; DF = detection frequency; DWTP = drinking water treatment plant; km = kilometer;
ND = not detected; ng/L = nanogram per liter; PFAA = perfluoroalkyl acid; PFAS = per- and polyfluoroalkyl substances; NR =
not reported; WWTP = wastewater treatment plant.
a The DF and/or mean was calculated using point data. Means were calculated only when DF = 100%.

Table B-3. Compilation of Studies Describing PFBS Occurrence in Surface Water

Study

Location

Site Details

PFBS Results

North America

Yeung et al.
(2017)

Canada (Ontario;
Mimico Creek, Rouge
River)

Two water samples at
each of the sites

Mimico Creek: point = 0.020
ng/L

Rouge River: DF 0%

79


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Study

Location

Site Details

PFBS Results

Subedi et al.
(2015)

United States (New
York; Skaneateles Lake)

Lake water along the
shoreline of residences
that use an enhanced
treatment unit for onsite
wastewater treatment

DFa 4% (n=28); single
detection value = 0.26 ng/L

Appleman et al.
(2014)

United States
(Wisconsin, Oklahoma,
Alaska, California,
Alabama, Colorado,
Ohio, Nevada,
Minnesota, New Jersey)

Raw surface waters from
11 sites, some impacted
by upstream wastewater
effluent discharge

DFa 64% (n=25); range = ND -
47 ng/L

(MRL = 0.3)

Veillette et al.
(2012)

Canada (Ellesmere
Island, Nunavut)

A lake near the northwest
coast with no known
sources of PFAS

DFa 100%, mean (range) =
0.016 (0.011-0.024) ng/L

Nakayama et al.
(2010)

United States (Illinois,
Iowa, Minnesota,
Missouri, Wisconsin;
Upper Mississippi River
Basin and Missouri
River Basin)

88 sampling sites from
tributaries and streams

DF 43%, median (range) = 0.71
(ND-84.1) ng/L

Galloway et al.
(2020)

United States (Ohio and
West Virginia; Ohio
River Basin)

Rivers and tributaries 58
km upstream to 130 km
downwind of a
fluoropolymer production
facility, some sample
locations potentially
impacted by local
landfills

DF NR, range3 = ND-28.0 ng/L

Newsted et al.
(2017)

United States
(Minnesota; Upper
Mississippi River Pool

2)

Upstream and
downstream of 3M
Cottage Grove facility
outfall, which is a source
of PFAS

Upstream: DFa 3%, point = 4.2
ng/L

Downstream: DFa 67%, range =
ND-336.0 ng/L

Procopio et al.
(2017)

United States (New
Jersey; Metedeconk
River Watershed)

Downstream of suspected
illicit discharge to soil
and groundwater from a
manufacturer of
industrial fabrics,
composites, and
elastomers that use or
produce products
containing PFAAs

DFa 5%, range = ND-100 ng/L

Newton et al.
(2017)

United States (Decatur,
Alabama; Tennessee
River)

6 sites upstream and 3
sites downstream of
fluorochemical
manufacturing facilities

Upstream: DF 0%

Downstream: DFa 100%, mean3
(range) = 69 (10-160) ng/L

80


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Study

Location

Site Details

PFBS Results

Zhang et al.
(2016)

United States (Rhode
Island, New York
Metropolitan Region)

Rivers and creeks, some
sampling locations
downstream from
industrial activities,
airport, textile mills, and
WWTP. PFAS are used
for water resistant
coating in textiles.

DFa 85%, range =ND-6.181
ng/L

Lescord et al.
(2015)

Canada (Resolute Bay,
Nunavut)

One lake (Meretta)
contaminated with runoff
from an airport, which is
a known source of PFAS;
one control lake (9 Mile)

Meretta: DF NR, mean = 4.9
ng/L

9 Mile: DF NR, mean = 0.07
ng/L

Lasier et al.
(2011)

United States (Georgia;
Coosa River watershed)

Upstream (sites 1 and 2)
and downstream (sites 3-
8) of a 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.

Upstream

Sites 1 and 2: DF 0%
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

Anderson et al.
(2016)

United States (national)

Ten U.S. Air Force
installations with historic
AFFF release

DF 80.00%, median (range) =
106 (ND-317,000) ng/L

Post et al.
(2013)

United States (New
Jersey)

6 rivers and 6 reservoirs
from public drinking
water system intakes,
some sites may include
nearby small industrial
park and civil-military
airport

DF 17%, range = ND-6 ng/L

Nakayama et al.
(2007)

United States (North
Carolina; Cape Fear
River Basin)

80 sampling sites in river
basin; some sites near
industrial areas and Fort
Bragg and Pope Air
Force Base with
suspected use of AFFF at
the Air Force Base

DF 62%, mean (range) = 2.58
(ND-9.41) ng/L

81


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Study

Location

Site Details

PFBS Results

Lindstrom et al.
(2011)

United States (Alabama)

32 surface water samples
(ponds and streams) from
areas with historical land
application of
fluorochemical industry-
impacted biosolids

DFa 63%, range = ND-208
ng/L

Bradley et al.
(2020)

United States (Lake
Michigan)

Untreated Lake Michigan
water from treatment
plant intake (4 sites)

DF 29%, range = ND-0.5 ng/L

Europe

Barreca et al.
(2020)

Italy (Lombardia
Region)

Rivers and streams with
no known fluorochemical
sources

DFa 39%, range = ND-16,000
ng/L

Loos et al.
(2017)

Austria, Bulgaria,
Croatia, Moldova,
Romania, Serbia,
Slovakia (Danube River
and tributaries)

Some sampling locations
downstream of major
cities

DF 94%, mean (range) = 1.6
(ND-3.7) ng/L

Wilkinson et al.
(2017)

England (Greater
London and southern
England; Hogsmill
River, Chertsey Bourne
River, Blackwater
River)

50 m upstream and 250
m and 1,000 m
downstream from WWTP
effluent outfalls

Upstream: DF NR, mean = 20.4
ng/L

Downstream 250 m: DF NR,
mean = 40.3 ng/L

Downstream 1,000 m: DF NR,
mean = 41.1 ng/L

Shafique et al.
(2017)

Germany (Leipzig,
PleiBe-Elster River,
Saale River, and Elbe
River)

Sampling sites were not
proximate to known point
sources of any
fluorochemical facilities

PleiBe-Elster: DF NR, mean =
1.2 ng/L

Saale: DF NR, mean = 7.5 ng/L
Elbe: DF NR, mean = 4.3 ng/L

Munoz et al.
(2016)

France (Seine River)

Two sites downstream of
Greater Paris and one site
unaffected by the Greater
Paris region

DF 70%, range = ND-3.1 ng/L

Lorenzo et al.
(2015)

Spain (Guadalquivir
River Basin, Ebro River
Basin)

Guadalquivir sampling
locations included
downstream ofWWTPs,
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

Guadalquivir: DF 8%, mean
(range) = 10.1 (ND-228.3)
ng/L

Ebro: DF 0%

82


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Study

Location

Site Details

PFBS Results

Zhao et al.
(2015)

Germany (Elbe River
and lower Weser River)

Some sampling sites near
Hamburg city and
industrial plants

Elbe: DF 100%, mean (range) =
7.4 (0.24-238) ng/L

Weser: DF 100%, mean (range)
= 1.41 (0.75-1.85) ng/L

Eriksson et al.
(2013)

Denmark (Faroe
Islands)

Lakes Leitisvatn,
Havnardal, Kornvatn, and
A Myranar with no
known point sources of
any fluorochemical
facilities

Leitisvatn: DF 0%
Havnardal Lake: DF 0%
Kornvatn Lake: DF 0%
A Myranar: DF 0%

Wagner et al.
(2013)

Germany (Rhine River)

Sampling sites were not
proximate to known point
sources of any
fluorochemical facilities

DFa 100%, meanb (rangeb) =18
(9-26) ng/L

Boiteux et al.
(2012)

France (national)

Rivers; some locations
may have upstream
industrial sources

DF 1%, range = ND-5 ng/L

Eschauzier et al.
(2012)

The Netherlands
(Amsterdam; Lek Canal,
tributary of Rhine River)

Downstream of an
industrial point source in
the German part of the
Lower Rhine

DFa 100%, mean (range) = 35
(31-42) ng/L

Labadie and

Chevreuil

(2011)

France (Paris; River
Seine)

Urban stretch of the
River Seine during a
flood cycle, sampling
location under the
influence of two urban
WWTPs and two major
combined sewer overflow
outfalls

DF 100%, mean (range) = 1.3
(0.6-2.6) ng/L

Moller et al.
(2010)

Germany (Rhine River
watershed)

Upstream and
downstream of
Leverkusen, where
effluent of a WWTP
treating industrial
wastewater was
discharged; other major
rivers and tributaries

Rhine upstream Leverkusen:
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

83


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Study

Location

Site Details

PFBS Results

Ahrens et al.
(2009b)

Germany (Elbe River)

Sampling sites in
Hamburg city (sites 16-
18) and from Laurenburg
to Hamburg (sites 19-24)

Hamburg:

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.
(2009a)

Germany (Elbe River)

Sampling locations 53 to
122 km (sites 1 to 9)c
upstream of estuary
mouth of Elbe River

DF NR; range of mean (for
different locations) = 1.8-3.4
ng/L

Rostkowski et
al. (2009)

Poland (national)

Rivers, lakes, and
streams in northern and
southern Poland, some
southern locations near
chemical industrial
activities

North: DFa 60%, range = ND-
10 ng/L

South: DFa 73%, range = ND-
16.0 ng/L

Ericson et al.
(2008b)

Spain (Tarragona
Province; Ebro River,
Francoli River, Cortiella
River)

Sampling sites were not
proximate to known point
sources of any
fluorochemical facilities

Ebro site 1: DF 0%
Ebro site 2: DF 0%
Francoli: DF 0%
Cortiella: DF 0%

Bach et al.
(2017)

France (southern)

Upstream and
downstream from
discharge point that
receives wastewater from
an industrial site with two
fluoropolymer
manufacturing facilities

Upstream: DF 0%
Downstream: DF 0%

Boiteux et al.
(2017)

France (northern)

River samples from
upstream and
downstream of an
industrial WWTP that
processes raw sewage
from fluorochemical
manufacturing facility

Upstream: DF 0%
Downstream: DF 0%

Gebbink et al.
(2017)

The Netherlands
(Dordrecht)

Upstream and
downstream of Dordrecht
fluorochemical
production plant; two
control sites

Control sites: DFa 100%, 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

84


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Study

Location

Site Details

PFBS Results

Valsecchi et al.
(2015)

Italy (Po River Basin,
Brenta River Basin,
Adige River Basin,
Tevere River Basin, and
Arno River Basin)

Two river basins (Po and
Brenta) which receive
discharges from
two chemical plants that
produce fluorinated
polymers and
intermediates; three river
basins (Adige, Tevere,
Arno) with no known
point sources of any
fluorochemical facilities

Po: DFa 56%, range = ND-30.4
ng/L

Brenta: DFa 100%, mean3
(range) = 707 (23.1-1,666)
ng/L

Adige: DFa 20%, range = ND-
4.3 ng/L

Tevere: DF 0%

Arno: DFa 58%, range = ND-
31.4 ng/L

Mussabek et al.
(2019)

Sweden (Lulea)

Samples from lake and
pond near a firefighting
training facility at the
Norrbotten Air Force
Wing known to use
PFAS-containing AFFF

Lake: DF NR, mean = 200 ng/L

Pond: DF NR, mean = 150
ng/L

Gobelius et al.
(2018)

Sweden (national)

Sampling locations
selected 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

DFa 29%, range = ND-299
ng/L

Dauchy et al.
(2017)

France (unspecified)

Samples collected near 3
sites (B, C, D) impacted
by the use of firefighting
foams

Site B: DF 0%

Site C: DF 0%

Site D: DFa 30%, range = ND-
138 ng/L

Multiple Continents

Pan et al. (2018)

United States (Delaware
River)

Sampling sites were not
proximate to known point
sources of any
fluorochemical facilities

DFa 100%, mean (range) = 2.19
(0.52-4.20) ng/L

United Kingdom
(Thames River)

Sampling sites were not
proximate to known point
sources of any
fluorochemical facilities

DFa 100%, mean (range) = 5.06
(3.26-6.75) ng/L

Germany and the
Netherlands (Rhine
River)

Sampling sites were not
proximate to known point
sources of any
fluorochemical facilities

DFa 100%, mean (range) = 21.9
(0.46-146) ng/L

85


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Study

Location

Site Details

PFBS Results



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; (ig/L = microgram per liter.

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%.

bFor 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 mS/cm.

Table B-4. Com

lilation of Studies Describing PFBS Occurrence in Food

Study

Location and Source

Food Types

Results

North America

Schecter et al.
(2010)

United States (Texas)
Grocery stores

Dairy, fruits and
vegetables,
grain, meat,
seafood,
fats/other

Cod: DF NR, mean = 0.12 ng/g ww

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

Byrne et al.
(2017)

United States (Alaska)

Upstream/downstream of
former defense site (Suqi
River)

Seafood

Blackfish: DF 48%, range = ND-
59.2 ng/g ww

Highest concentration was upstream

Scher et al.
(2018)

United States (Minnesota)

Home gardens

Near former 3M PFAS
production facility, homes
within and outside a GCA

Fruits and
vegetables

Within GCA:

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

Blaine et al.
(2014)

United States (Midwestern)

Greenhouse study,
unamended controls

Fruits and
vegetables

Radish root: DF NR, mean = 22.36

ng/g

ND in celery shoot, pea fruit

86


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Study

Location and Source

Food Types

Results

Blaine et al.
(2013)

United States (Midwestern)

Greenhouse and field
studies, unamended controls

Fruits and

vegetables,

grain

ND in corn, lettuce, tomato in
unamended soil.

Young et al.
(2013)

United States (Maryland,
Mississippi, Tennessee,
Florida, New York, Texas,
Washington, D.C.)

Retail markets

Seafood

ND in crab, shrimp, striped bass,
farm raised catfish, farm raised
salmon

Young et al.
(2012)

United States (17 states)
Retail markets

Dairy

ND in retail cow's milk

Europe

Domingo et al.
(2012)

Spain (Catalonia)

Local markets, small stores,
supermarkets, big grocery
stores

12 food
categories

Vegetables: DF NR, 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

Perez et al.
(2014)

Serbia (Belgrade and Novi
Sad), Spain (Barcelona,
Girona, and Madrid)

Various supermarkets and
retail stores

8 food
categories

Categories included cereals, pulses
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)

D'Hollander et al.
(2015)

Belgium, Czech Republic,
Italy, Norway

PERFOOD study; items
from 3 national retail stores
of different brands and
countries of origin

Fruit, cereals,
sweets, salt

Sweets: DFa 25%, range = ND-
0.0016 ng/g

Fruit: DFa 19%, range = ND-
0.067 ng/g

ND in cereals, salt

Hlouskova et al.
(2013)

Belgium, Czech Republic,
Italy, Norway

Several national
supermarkets

Pooled
milk/dairy
products, meat,
fish, hen eggs

DF 5%, mean (range) = 0.00975
(0.006-0.012) ng/g

87


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Study

Location and Source

Food Types

Results

Eriksson et al.

Denmark

Dairy, fruits and

Milk:

(2013)

Farm, dairy farm, fish from
Faroe Shelf area

vegetables,
seafood

Farmer (Havnardal): point =
0.019 ng/g ww

Diary (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

Sznajder-

Poland

Dairy

All dairy: sum PFBS = 0.04 ng/g

Katarzynska et
al. (2019)

Markets



Butter: range = 0.01-0.02 ng/g

ND in camembert-type cheese,
cottage cheese, milk, natural yogurt,
sour cream, kefir (bonny clabber)

Yamada et al.

France

Seafood

Freshwater fish: DF NR, range =

(2014)

Freshwater fish from 6
major French rivers; fresh
and frozen fish from markets



0.06-0.16 ng/g ww

Fresh or frozen fish: DF NR,
range = 0.02-0.03 ng/g ww

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

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

88


-------
Study

Location and Source

Food Types

Results

Surma et al.
(2017)

Spain, Slovakia
Source NR

Fats/other

Spices: ND-1.01 ng/g

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

Papadopoulou et
al. (2017)

Norway

A-TEAM project: food and
drinks collected by
participants as duplicate diet
samples

Solid foods (11
food

categories),
liquid foods (5
drinks)

Solid foods (unspecific food
category): DF 2%, range = ND-
0.001 ng/g

ND in liquid foods (coffee, tea and
cocoa, milk, water, alcoholic
beverages and soft drinks)

Scordo et al.
(2020)

Italy

Supermarkets

Fruits

Olives: DFa 100%, mean" (range)
= 0.294 (0.185-0.403) ng/g dw

ND in strawberries

Ericson et al.
(2008a)

Spain

Local markets, large
supermarkets, grocery stores

18 food
categories

ND in all categories: veal, pork,
chicken, lamb, white fish, seafood,
tinned fish, blue fish, whole milk,
semi-skimmed milk, dairy products,
vegetables, pulses, cereals, fruits,
oil, margarine, and eggs

Noorlander et al.
(2011)

The Netherlands

Several Dutch retail store
chains with nationwide
coverage

15 food
categories

ND in all categories: flour, fatty fish,
lean fish, pork, eggs, crustaceans,
bakery products, vegetables/fruit,
cheese, beef, chicken/poultry, butter,
milk, vegetable oil, and industrial oil

Jogsten et al.
(2009)

Spain (Catalonia)

Local markets, large
supermarkets, grocery stores

Fruits and
vegetables,
meat, seafood,
fats/other

ND in lettuce, raw, cooked, and fried
meat (veal, pork, and chicken), fried
chicken nuggets, black pudding,
lamb liver, pate of pork liver, foie
gras of duck, "Frankfurt" sausages,
home-made marinated salmon, and
common salt

Sznajder-
Katarzynska et
al. (2018)

Poland
Markets

Fruits and
vegetables

ND in apples, bananas, cherries,
lemons, oranges, strawberries,
beetroots, carrots, tomatoes,
potatoes, and white cabbage

89


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Study

Location and Source

Food Types

Results

Falandysz et al.
(2006)

Poland

Gulf of Gdansk, Baltic Sea
south coast

Meat, seafood

ND in eider duck, cod

Barbosa et al.
(2018)

Belgium, France, the
Netherlands, Portugal

Various markets

Seafood

ND in raw and steamed fish (P.

platessa, M. australis, M. capenis,
K. pelamis, and M. edulis)

Holzer et al.
(2011)

Germany

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

Seafood

Lake Mohne /River Mohne: ND in
cisco, eel, perch, pike, and roach

Trade/markets: ND in eel,
pike/perch, and trout

Jorundsdottir et
al. (2014)

Iceland

Collected during biannual
scientific surveys,
commercially-produced

Seafood

ND in anglerfish, Atlantic cod, blue
whiting, lemon sole, ling, lumpfish,
plaice, and pollock

Riviere et al.
(2019)

France

Based on results of national
consumption survey

Seafood,
fats/other

ND in infant food, vegetables, non-
alcoholic beverages, dairy-based
desserts, milk, mixed dishes, fish,
ultra-fresh dairy products, meat,
poultry and game

Lankova etal.
(2013)

Czech Republic
Retail market

Fats/other

ND in infant formula

Zafeiraki et al.
(2016a)

Greece, the Netherlands

Home and commercially-
produced

Fats/other

ND in chicken eggs

Gebbink et al.
(2015)

Sweden

Major grocery chain stores,
market basket samples

12 food
categories

ND in all categories: dairy products,
meat products, fats, pastries, fish
products, egg, cereal products,
vegetables, fruit, potatoes, sugar and
sweets, soft drinks

Herzke et al.
(2013)

Belgium, Czech Republic,
Italy, Norway

PERFOOD study: items
from 3 national retail stores
of different brands per
location

Vegetables

ND for all vegetables

Zafeiraki et al.
(2016b)

The Netherlands

Local markets and
slaughterhouses

Meat

ND for horse, sheep, cow, pig, and
chicken liver

90


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Study

Location and Source

Food Types

Results

Multiple Continents

Chiesa etal.
(2019)

United States (Pacific
Ocean)

Wholesale fish market

Seafood

ND in wild-caught salmon

Canada

Wholesale fish market

Seafood

ND in wild-caught salmon

Norway

Wholesale fish market

Seafood

ND in farm salmon

Scotland

Wholesale fish market

Seafood

ND in wild-caught and farm salmon

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; (ig/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%.

Table B-5. Compilation of Studies Describing PFBS Occurrence in Indoor Dust

Study

Location

Site Details

Results

North America

Zheng et al. (2020)

United States
(Seattle, Washington
and West Lafayette,
Indiana)

Childcare facilities (20
samples from 7 in
Seattle and 1 in West
Lafayette)

DF 90%, mean (range) = 0.34
(ND-0.86) ng/g

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. (2013)

United States

(Boston,

Massachusetts)

Homes (30); offices
(31); vehicles (13)

Homes: DF 3% (single detection),
range = ND-4.98 ng/g

Offices: DF 10%, range = ND-
12.0 ng/g

Vehicles: DF 0%

Knobeloch et al.
(2012)

United States (Great
Lakes Basin,
Wisconsin)

Homes (39)

DF 59%, median (range) = 1.8
(ND-31) ng/g

91


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Study

Location

Site Details

Results

Strynar and
Lindstrom (2008)

United States (Cities
in North Carolina and
Ohio)

Homes (102) and
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

DF 33%, mean (range) = 41.7
(ND-1,150) ng/g

Scheretal. (2019)

United States (Twin
Cities metropolitan
region, Minnesota)

Near former 3M PFAS
production facility; 19
homes within the
GCA

Entryway: DF 11%, median
(range) = ND (ND-58 ng/g)

Living room: DF 16%, median
(range) = ND (ND-58 ng/g)

Kubwabo et al.
(2005)

Canada (Ottawa)

Homes (67)

DF 0%

Europe

de la Torre et al.
(2019)

Spain (unspecified),
Belgium

(unspecified), Italy
(unspecified)

Homes (65)

Spain: DF 52%, median (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

Harrad etal. (2019)

Ireland (Dublin,
Galway, and
Limerick counties)

Homes (32); offices
(33); cars (31);
classrooms (32)

Homes: DF 81%, mean (range) =
17 (ND-110) ng/g

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

Giovanoulis et al.
(2019)

Sweden (Stockholm)

Preschools (20)

DF 0%

Winkens et al.
(2018)

Finland (Kuopio)

Homes (63 children's
bedrooms)

DF 12.7%, median (range) = ND
(ND-13.5) ng/g

Padilla-Sanchez
and Haug (2016)

Norway (Oslo)

Homes (7)

DF 14% (single detection), range
= ND-3 ng/g

Jogsten et al.
(2012)

Spain (Catalonia)

Homes (10)

DF 60%, range = ND-6.5 ng/g

Haug etal. (2011)

Norway (Oslo)

Homes (41)

DF 22%, mean (range) = 1.3
(0.17-9.8) ng/g

92


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Study

Location

Site Details

Results

Huber etal. (2011)

Norway (Tromso)

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

D'Hollander et al.
(2010)

Belgium (Flanders)

Homes (45); offices
(10)

Homes: DF 47%, median = 0 ng/g
dw

Offices: DF NR, median = 0.2
ng/g dw

Multiple Continents

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

Karaskova et al.
(2016)

United States
(unspecified)

Homes (14)

DF 60%, mean (range) = 1.4 (ND-
2.6) ng/g



Canada (unspecified)

Homes (15)

DF 55%, mean (range) = 1.6 (ND-
5.8) ng/g



Czech Republic
(unspecified)

Homes (12)

DF 37.5%, mean (range) = 3.6
(ND-14.4) ng/g

Notes: DF = detection frequency; GCA = groundwater contamination area; ND = not detected; ng/g = nanogram per gram; NR =
not reported; dw = dry weight

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%.

Table B-6. Compilation of Studies Describing PFBS Occurrence in Soil

Study

Location

Site Details

Results

North America

Venkatesan and
Halden (2014)

United States

(Baltimore,

Maryland)

Control (nonamended)
soil from Beltsville
Agricultural Research
Center

DF 0%

93


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Study

Location

Site Details

Results

Blaine et al. (2013)

United States
(Midwestern)

Urban and rural full-
scale field study
control (nonamended)
soil

Urban control: DF NR, mean =
0.10 ng/g

Rural control: DF NR, mean = ND

Scheretal. (2019)

United States (Twin
Cities metropolitan
region, Minnesota)

Near former 3M PFAS
production facility,
homes within a GCA

DF 10%, median (p90) = ND
(0.02) ng/g

Scheretal. (2018)

United States (Twin
Cities metropolitan
region, Minnesota)

Near former 3M PFAS
production facility,
homes within and
outside a GCA

Within GCA: DF 9%, median
(range) = ND (ND-0.17 ng/g)

Outside GCA: DF 17%, median
(range) =ND (ND-0.031 ng/g)

Anderson et al.
(2016)

United States
(unspecified)

Ten U.S. Air Force
installations with
historic AFFF release,
surface and subsurface
soils

Surface soil: DF 35%, median
(range) = 0.775 (ND-52.0) ng/g

Subsurface soil: DF 35%, median
(range) = 1.30 (ND-79.0) ng/g

Eberle et al. (2017)

United States (Joint
Base Langley-
Eustis, Virginia)

Firefighting training
site, pre- and
posttreatment

Pretreatment: DF 60%, range =
0.61-6.4 ng/g

Posttreatment: DF 100%, range =
0.07-0.83 ng/g

Nickerson et al.
(2020)

United States
(unspecified)

Two AFFF-impacted
soil cores from former
fire-training areas

Core E: DFa 91%, range = ND-
27.37 ng/g dw

Core F: DF 100%, range = 0.13—
58.44 ng/g dw

Cabrerizo et al.
(2018)

Canada (Melville
and Cornwallis
Islands)

Catchment areas of
lakes

DF 100%, mean3 (range) = 0.0024
(0.0004-0.0083) ng/g dw

Dreyer et al. (2012)

Canada (Ottawa,
Ontario)

Mer Bleue Bog Peat
samples (core samples)

Detected once at 0.071 ng/g in
1973 sample and not considered
for further evaluation

Mejia-Avendano et
al. (2017)

Canada (Lac-
Megantic, Quebec)

Site of 2013 Lac-
Megantic train accident
(oil and AFFF runoff
area [sampled 2013],
burn site and adjacent
area [sampled 2015])

Background: DF NR, 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

Europe

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

94


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Study

Location

Site Details

Results

Gronncstad 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)

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%

Dauchy et al.
(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

Skaaretal. (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

Hale et al. (2017)

Norway
(Gardermoen)

Firefighting training
site

DF 0%

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
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%.

b Grannestad et al. (2019) reported a DF = 10% but a range, mean, and standard deviation of < LOQ.

95


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