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


EPA/690/R-21/001F | April 2021 | www.epa.gov/pprtv
United States
Environmental Protection
Agency
-ivEPA
Provisional Peer-Reviewed Toxicity Values for
Perfluorobutane Sulfonic Acid
(CASRN 375-73-5)
and Related Compound
Potassium Perfluorobutane Sulfonate
(CASRN 29420-49-3)
U.S. EPA Office of Research and Development
Center for Public Health and Environmental Assessment

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EPA/690/R-21/001F
April 2021
www.epa.gov/pprtv
Provisional Peer-Reviewed Toxicity Values for
Perfluorobutane Sulfonic Acid
(CASRN 375-73-5)
and Related Compound
Potassium Perfluorobutane Sulfonate
(CASRN 29420-49-3)
Center for Public Health and Environmental Assessment
Office of Research and Development
U.S. Environmental Protection Agency
Cincinnati, OH 45268

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EPA/690/R-21/001F
DISCLAIMER
This document has been reviewed in accordance with U.S. Environmental Protection
Agency policy and approved for publication. Mention of trade names or commercial products
does not constitute endorsement or recommendation for use.
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EPA/690/R-21/001F
AUTHORS, CONTRIBUTORS, AND REVIEWERS
CHEMICAL MANAGERS
Jason C. Lambert, PhD, DABT
Elizabeth Oesterling Owens, PhD
CONTRIBUTORS
Michelle Angrish, PhD
Xabier Arzuaga, PhD
Johanna Congleton, PhD
Ingrid Druwe, PhD
J. Allen Davis, MS
Kelly Garcia, BS
Carolyn Gigot, BA
Andrew Greenhalgh, BS
Joshua Harrill, PhD
Belinda Hawkins, PhD, DABT
Ryan Jones, MSLS
Andrew Kraft, PhD
Yu-Sheng Lin, PhD
April Luke, PhD
Elizabeth Radke, PhD
Paul Schlosser, PhD
Michele Taylor, PhD
Samuel Thacker, BA
Andre Weaver, PhD
Jacqueline Weinberger, BS
Amina Wilkins, MPH
Michael Wright, ScD
Q. Jay Zhao, PhD, MPH, DABT
DRAFT DOCUMENT PREPARED BY
U.S. Environmental Protection Agency, Office of Research and Development,
Center for Public Health and Environmental Assessment (CPHEA)
EXECUTIVE DIRECTION
Wayne Cascio, MD, FACC
V. Kay Holt
Samantha Jones, PhD
Emma Lavoie, PhD
Tina Bahadori, ScD
Mary Ross, PhD
CPHEA Center Director
CPHEA Deputy Center Director
CPHEA Associate Director
CPHEA Senior Advisor
NCEA Center Director (previous Exec. Direction)
NCEA Deputy Center Director (previous Exec. Direction)
PRIMARY INTERNAL REVIEWERS
David Bussard
Kristina Thayer, PhD
Scott Wesselkamper, PhD
IV

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EPA/690/R-21/001F
PREFACE
This assessment titled Provisional Peer-Reviewed Toxicity Values for Perfluorobutane
Sulfonic Acid and Related Compound Potassium Perfluorobutane Sulfonate is a toxicity
assessment developed by the U.S. EPA's Office of Research and Development (ORD) Center for
Public Health and Environmental Assessment (CPHEA).
A Provisional Peer-Reviewed Toxicity Value (PPRTV) is defined as a toxicity value
derived for use in the Superfund Program. This assessment for perfluorobutane sulfonic acid
(PFBS) updates and replaces the 2014 PPRTV assessment for PFBS. Currently available
PPRTV assessments can be accessed on the U.S. Environmental Protection Agency's
(U.S. EPA's) PPRTV website at https://www.epa.gov/pprtv. Questions regarding nomination of
chemicals for update can be sent to the appropriate U.S. EPA Superfund and Technology Liaison
(https://www.epa.gov/research/fact-sheets-regional-science). This assessment is also available
for use across multiple U.S. EPA program and regional offices, other federal agencies, states,
tribes, external stakeholders, and other entities as needed as a Human Health Toxicity Value
Assessment.
The perfluorobutane sulfonic acid (PFBS) toxicity assessment is one of the key goals of
the Agency's PFAS Action Plan (U.S. EPA. 2019) and provides qualitative and quantitative
toxicity information that can be used along with exposure information and other important
considerations to assess potential health risks to determine if, and when, it is appropriate to take
action to address this chemical.
The PFBS human health toxicity values presented in this assessment were developed
based on the best available science. The assessment provides high-quality evaluations and
conclusions drawn from publicly available information on the toxicity of PFBS. This assessment
is not a regulation; rather, it provides a critical part of the scientific foundation for risk
assessment decision making. The PFBS assessment provides toxicity values and information
about the adverse effects of the chemical and the evidence on which the value is based, including
the strengths and limitations of the data. All users, including risk assessors and risk managers,
are advised to review the information, including potential uncertainties, provided in this
document to ensure that the assessment is appropriate for the circumstances (e.g., exposure
pathways, concentrations, presence of sensitive subpopulations) in question and the risk
management decisions that would be supported by the risk assessment.
The PFBS toxicity assessment underwent a rigorous development and review process, as
described below.
Overview of major steps in the PFBS assessment development and review process
• Draft assessment development
• Review by U.S. EPA program and regional offices (i.e., Intra-agency review)
• Review by other federal agencies (i.e., interagency review)
• External letter peer review
• Public comment period
• Second external letter peer review
• Intra-agency and interagency review
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EPA/690/R-21/001F
This assessment was provided for review to scientists in U.S. EPA's program and
regional offices prior to external peer review and after external peer review. Comments were
submitted by:
Office of the Administrator/Office of Children's Health Protection
Office of the Administrator/Office of Policy
Office of Chemical Safety and Pollution Prevention
Office of Land and Emergency Management
Office of Research and Development
Office of Water
Region 2, New York, NY
Region 3, Boston, MA
Region 4, Atlanta, GA
Region 5, Chicago, IL
Region 8, Denver, CO
This assessment was provided for review to other federal agencies prior to external peer
review and after external peer review. Representatives from federal agencies and from the
Environmental Council of the States (ECOS) were briefed during the assessment scoping and
draft development process on March 9, 2018; May 2, 2018; and August 27, 2018After public
comment, interagency review was conducted by the Office of Management and Budget's PFAS
Technical Working Group (TWG), an interagency group composed of career staff chief scientists
or their equivalents from across the Executive Branch. Comments on this assessment were
submitted by a subset of TWG representatives, namely:
Department of Defense
Department of Health and Human Services
Agency for Toxic Substances and Disease Registry
Food and Drug Administration
National Institute of Environmental Health Sciences/National Toxicology Program
National Institute of Occupational Safety and Health
Executive Office of the President
Office of Management and Budget
National Aeronautics and Space Administration
This assessment was peer reviewed by independent, expert scientists external to
U.S. EPA before and after the public comment period. The reports of the two external peer
reviews and responses to comments on the U.S. EPA's draft Human Health Toxicity Values for
PFBS, dated November 2018 and October 2020, are available at https://www.epa.gov/pfas/learn-
about-human-health-toxicitv-assessment-pfbs. Comments from external peer review were
submitted by:
Karen Chou, PhD
Dale Hattis, PhD
Lisa M. Kamendulis, PhD
Angela M. Leung, MD
Angela L. Slitt, PhD
David Alan Warren, MPH, PhD
R. Thomas Zoeller, PhD
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This assessment was released for public comment from November 21, 2018 to January
22, 2019. The public comments are available on Regulations.gov in the Docket ID
No. EPA-HQ-OW-2018-0614.
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TABLE OF CONTENTS
Disclaimer	iii
Authors, Contributors, and Reviewers	iv
Preface	v
List of Figures	x
List of Tables	xii
Commonly Used Abbreviations and Acronyms	xiii
Executive Summary	1
1.0 Background	5
1.1	Physical and Chemical Properties	5
1.2	Occurrence	6
1.3	Toxicokinetics	8
1.3.1	Overview	8
1.3.2	Absorption	11
1.3.3	Distribution	11
1.3.4	Metabolism	13
1.3.5	Excretion	13
1.3.6	Physiologically Based Pharmacokinetic Models	16
1.3.7	Summary	17
2.0 Problem Formulation	19
2.1	Conceptual Model	19
2.2	Objective	21
2.3	Methods	21
2.3.1	Literature Search	21
2.3.2	Screening Process	21
2.3.3	Study Evaluation	22
2.3.4	Data Extraction	24
2.3.5	Evi dence Synthe si s	25
2.3.6	Evidence Integration and Hazard Characterization	25
2.3.7	Derivation of Values	27
3.0 Overview of Evidence Identification for Synthesis and Dose-Response Analysis	30
3.1	Literature Search and Screening Results	30
3.2	Study Evaluation Results	31
4.0 Evidence Synthesis: Overview of Included Studies	35
4.1	Thyroid Effects	35
4.1.1	Human Studies	35
4.1.2	Animal Studies	36
4.2	Reproductive Effects	37
4.2.1	Human Studies	37
4.2.2	Animal Studies	37
4.3	Offspring Growth and Early Development	42
4.3.1	Human Studies	42
4.3.2	Animal Studies	42
4.4	Renal Effects	43
4.4.1	Human Studies	43
4.4.2	Animal Studies	43
4.5	Hepatic Effects	44
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4.5.1	Human Studies	44
4.5.2	Animal Studies	44
4.6	Effects on Lipids or Lipoproteins	45
4.6.1	Human Studies	45
4.6.2	Animal Studies	46
4.7	Other Effects	47
4.7.1	Human Studies	47
4.7.2	Animal Studies	48
4.8	Other Data	48
4.8.1	Tests Evaluating Genotoxicity and Mutagenicity	52
4.8.2	Acute Duration and Other Routes of Exposure	52
5.0 Evidence Integration and Hazard Characterization	53
5.1	Thyroid Effects	56
5.2	Developmental Effects	57
5.3	Reproductive Effects	57
5.4	Renal Effects	59
5.5	Hepatic Effects	60
5.6	Effects on Lipids or Lipoproteins	60
5.7	Immune Effects	61
5.8	Cardiovascular Effects	61
5.9	Evidence Integration and Hazard Characterization Summary	61
6.0 Derivation of Values	73
6.1	Derivation of Oral Reference Doses	73
6.1.1	Derivation of the Subchronic Oral Reference Dose	73
6.1.2	Derivation of the Chronic Oral Reference Dose	86
6.2	Derivation of Inhalation Reference Concentrations	89
6.3	Cancer Weight-of-Evidence Descriptor and Derivation of Cancer Risk Values	89
6.4	Susceptible Populations and Life Stages	90
Appendix A. Literature Search Strategy	91
Appendix B. Detailed PECO Criteria	93
Appendix C. Study Evaluation Methods	94
Appendix D. HAWC User Guide and Frequently Asked Questions	Ill
Appendix E. Additional Data Figures	118
Appendix F. Benchmark Dose Modeling Results	127
Appendix G. Quality Assurance	141
Appendix H. References	143
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LIST OF FIGURES
Figure 1. PFBS andK+PFBS Chemical Structures	5
Figure 2. Conceptual Model for PFBS and/or Potassium Salt	20
Figure 3. Approach for Evaluating Epidemiological and Animal Toxicology Studies	24
Figure 4. Literature Search and Screening Flow Diagram for PFBS (CASRN 375-73-5)	31
Figure 5. Evaluation Results for Epidemiological Studies Assessing Effects of PFBS
(Click to see interactive data graphic for rating rationales)	33
Figure 6. Evaluation Results for Animal Studies Assessing Effects of PFBS Exposure
(Click to see interactive data graphic for rating rationales)	34
Figure 7. Thyroid Effects from K+PFBS Exposure (Click to see interactive data graphic
and rationale for study evaluations for effects on the thyroid in HAWC)	36
Figure 8. Reproductive Hormone Response to K+PFBS Exposure (Click to see interactive
data graphic and rationale for study evaluations for reproductive hormone
levels in HAWC)	39
Figure 9. Effects on Reproductive Development and Estrous Cycling Following
PFBS Exposure (Click to see interactive data graphic)	41
Figure D-l. HAWC Homepage for the Public PFBS Assessment	Ill
Figure D-2. Representative Study List	112
Figure D-3. Representative Study Evaluation Pie Chart with the Reporting Domain
Selected and Text Populating to the Right of Pie Chart	113
Figure D-4A. Visualization Example for PFBS (Note that the records listed under each
column [study, experiment endpoint, units, study design, observation time,
dose] and data within the plot are interactive.)	114
Figure D-4B. Example Pop-Up Window after Clicking on Interactive Visualization Links
(In Figure D-4A, the red circle for study NTP (2019); male at a dose of
500 mg/kg-day was clicked leading to the pop-up shown above. Clicking
on the blue text will open a new window with descriptive data.)	115
Figure D-5. Representative Data Download Page	116
Figure D-6A. Example BMD Modeling Navigation	117
Figure D-6B. Example BMD Session	117
Figure E-l. Serum Free and Total Thyroxine (T4) Response in Animals Following
K+PFBS Exposure (Click to see interactive data graphic)	118
Figure E-2. Serum Total Triiodothyronine (T3) Response in Animals Following
K+PFBS Exposure (Click to see interactive data graphic)	119
Figure E-3. Serum Thyroid-Stimulating Hormone (TSH) Response in Animals Following
K+PFBS Exposure (Click to see interactive data graphic)	120
Figure E-4. Developmental Effects (Eye Opening) Following K+PFBS Exposure in Rats
(Click to see interactive data graphic)	120
Figure E-5. Developmental Effects (First Estrus) Following K+PFBS Exposure in Rats
(Click to see interactive data graphic)	121
Figure E-6. Developmental Effects (Vaginal Patency) Following K+PFBS Exposure in
Rats (Click to see interactive data graphic)	121
Figure E-7. Kidney Histopathological Effects Following K+PFBS Exposure in Rats
(Click to see interactive data graphic)	122
Figure E-8. Renal Effects Following K+PFBS Exposure in Rats (Click to see interactive
data graphic)	123
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Figure E-9. Kidney-Weight Effects Following K+PFBS Exposure in Rats (Click to see
interactive data graphic)	124
Figure E-10. Liver Effects Following K+PFBS Exposure in Rats (Click to see interactive
data graphic)	125
Figure E-l 1. Effects on Lipids and Lipoproteins Following K+PFBS Exposure in Rats
and Mice (Click to see interactive data graphic)	126
Figure F-l. Candidate PODs for the Derivation of the Subchronic and Chronic RfDs for
PFBS (Click to see interactive data graphic)	129
Figure F-2. Exponential (Model 4) for Total T4 in PND 1 Female Offspring (Litter n)
Exposed GDs 1-20 (Feng et al., 2017)	131
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LIST OF TABLES
Table 1. Physicochemical Properties ofPFBS (CASRN 375-73-5) and Related
Compound K+PFBS (CASRN 29420-49-3)	6
Table 2. Summary of the Toxicokinetics of Serum PFBS (Mean ± SE)	9
Table 3. Criteria for Overall Evidence Integration Judgments	26
Table 4. Epidemiological Studies Excluded Based on Study Evaluation	32
Table 5. Other Studies	49
Table 6. Summary of Noncancer Data for Oral Exposure to PFBS (CASRN 375-73-5)
and the Related Compound K PFBS (CASRN 29420-49-3)	54
Table 7. Summary of Hazard Characterization and Evidence Integration Judgments	62
Table 8. Mouse, Rat, and Human Half-Lives and Data-Informed DAFs	79
Table 9. PODs Considered for Deriving the Subchronic RfD for
K PFBS (CASRN 29420-49-3)	79
Table 10. Uncertainty Factors for the Subchronic RfD for Thyroid Effects for
K PFBS (CASRN 29420-49-3)	84
Table 11. Confidence Descriptors for the Subchronic RfD for PFBS (CASRN 375-73-5)
and the Related Compound K PFBS (CASRN 29420-49-3)	86
Table 12. Uncertainty Factors for the Chronic RfD for Thyroid for
K PFBS (CASRN 29420-49-3)	87
Table 13. Confidence Descriptors for Chronic RfD for PFBS (CASRN 375-73-5) and the
Related Compound K PFBS (CASRN 29420-49-3)	89
Table A-l. Synonyms and MeSH Terms	91
Table B-l. Population, Exposure, Comparator, and Outcome Criteria	93
Table C-l. Questions Used to Guide the Development of Criteria for Each Domain in
Epidemiology Studies	95
Table C-2. Criteria for Evaluation of Exposure Measurement in Epidemiology Studies	98
Table C-3. Questions Used to Guide the Development of Criteria for Each Domain in
Experimental Animal Toxicology Studies	101
Table F-l. Candidate PODs for the Derivation of the Subchronic and Chronic RfDs for
PFBS (CASRN 375-73-5) and the Related Compound K+PFBS
(CASRN 29420-49-3)	127
Table F-2. Modeling Results for Total T4 in PND 1 Female Offspring (Litter n)
Exposed GDs 1 20a	131
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COMMONLY USED ABBREVIATIONS AND ACRONYMS
AEC	absolute eosinophil count
AFFF	aqueous film-forming foam
AIC	Akaike's information criterion
ALT	alanine aminotransferase
AST	aspartate aminotransferase
AUC	area under the curve
BMD	benchmark dose
BMDL	benchmark dose lower confidence
limit
BMDS	Benchmark Dose Software
BMR	benchmark response
BUN	blood urea nitrogen
BW	body weight
CA	chromosomal aberration
CASRN	Chemical Abstracts Service registry
number
CHO	Chinese hamster ovary (cell line)
CI	confidence interval
CPHEA	Center for Public Health and
Environmental Assessment
CPN	chronic progressive nephropathy
D3	deiodinase 3
DAF	dosimetric adjustment factor
DNA	deoxyribonucleic acid
ECP	eosinophilic cationic protein
GD	gestation day
GLP	Good Laboratory Practice
HAWC	Health Assessment Workspace
Collaborative
HED	human equivalent dose
HPT	hypothalamic-pituitary-thyroid
i.v.	intravenous
ICR	Institute of Cancer Research
K+PFBS	potassium perfluorobutane sulfonate
keiim	serum elimination rate constant
LD	lactation day
LD50	median lethal dose
LOAEL	lowest-observed-adverse-effect
level
MW	molecular weight
NHANES National Health and Nutrition
Examination Survey
NOAEL	no-observed-adverse-effect level
NTP	National Toxicology Program
NZW	New Zealand White (rabbit breed)
OR	odds ratio
PECO	Population, Exposure, Comparator, and
Outcome
PFAA	perfluoroalkyl acid
PFAS	per- and polyfluoroalkyl substances
PFBS	perfluorobutane sulfonic acid
PFHxA	perfluorohexanoic acid
PFOA	perfluorooctanoic acid
PFOS	perfluorooctane sulfonic acid
PND	postnatal day
POD	point of departure
RfC	inhalation reference concentration
RfD	oral reference dose
ROS	reactive oxygen species
rT3	reverse triiodothyronine
S-D	Sprague-Dawley
SD	standard deviation
T2	3,5-diiodo-L-thyronine
T3	triiodothyronine
T4	thyroxine
TBG	thyroid-binding globulin
TSH	thyroid-stimulating hormone
TTR	transthyretin
UF	uncertainty factor
UFa	interspecies uncertainty factor
UFc	composite uncertainty factor
UFd	database uncertainty factor
UFh	intraspecies uncertainty factor
UFl	LOAEL-to-NOAEL uncertainty factor
UFs	subchronic-to-chronic uncertainty
factor
U.S. EPA	U.S. Environmental Protection Agency
VLDL	very low-density lipoprotein
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EXECUTIVE SUMMARY
SUMMARY OF OCCURRENCE AND HEALTH EFFECTS
The U.S. Environmental Protection Agency (U.S. EPA) is issuing subchronic and chronic
oral toxicity values for perfluorobutane sulfonic acid (PFBS) (Chemical Abstracts Service
registry number [CASRN] 375-73-5) and its related salt, potassium perfluorobutane sulfonate
(K+PFBS) (CASRN 29420-49-3). The ionic state of per- and polyfluoroalkyl substances (PFAS)
such as PFBS influence physicochemical properties such as water or lipid solubility and
bioaccumulative potential, which in turn impact fate and transport in the environment and
potential human health and ecological effects in exposed populations. K+PFBS fully dissociates
in aqueous solutions with pH levels ranging from 4-9; thus, the oral toxicity values derived in
this document are also applicable to the deprotonated anionic form of PFBS (i.e., PFBS ;
CASRN 45187-15-3).
The toxicity assessment for PFBS includes toxicity values associated with potential
noncancer health effects following oral exposure (in this case, oral reference doses [RfDs]). This
assessment evaluates human health hazards. The toxicity assessment and the values contained
within is not a risk assessment because it does not include an exposure assessment nor an overall
risk characterization. Further, the toxicity assessment does not address the legal, political, social,
economic, or technical considerations involved in risk management. The PFBS toxicity
assessment can be used by U.S. EPA, states, tribes, and local communities, along with specific
exposure and other relevant information, to determine, under the appropriate regulations and
statutes, if, and when, it is necessary to take action to address potential risk associated with
human exposures to PFBS.
PFBS and K+PFBS are both four-carbon, fully fluorinated alkane members of a large and
diverse class of linear and branched compounds known as "per- and polyfluoroalkyl substances,"
or PFAS. In the early 2000s, concerns grew over the environmental persistence,
bioaccumulation potential, and long half-lives in humans of longer chain PFAS, in particular,
perfluorooctanoic acid (PFOA) and perfluorooctane sulfonic acid (PFOS). As a result, shorter
chain PFAS such as PFBS were developed and integrated into various consumer products and
industrial applications, because PFBS has the desired properties and characteristics associated
with this class of compounds but with faster elimination from the body than PFOA and PFOS.
PFBS is associated with aqueous film-forming foam (foams (AFFFs) and used during chrome
electroplating as a mist suppressant (See Section 1.2). It has also been found in food contact
materials, dust, and source and finished drinking water. Accordingly, oral intake of water and
food, inhalation, and dermal contact are plausible modes of PFBS exposure, with the oral route
being the primary route of exposure. PFBS has been detected in humans, confirming exposure to
this PFAS; however, the magnitude of human exposure likely depends on factors such as
occupation (e.g., processing and/or manufacture of PFBS or PFBS-containing products and
chrome electroplating) and living conditions (e.g., proximity to locations that make or use
PFBS-containing products and nearby well-water use).
Human studies have examined possible associations between PFBS exposure and
potential health outcomes such as alteration of menstruation, reproductive hormones or semen
parameters, kidney function (uric acid production), lung function (induction of asthma), and lipid
profile. The ability to draw conclusions about associations is limited due to the small number of
human studies per outcome. Of the examined health outcomes, only asthma and serum
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cholesterol levels in humans were found to exhibit a statistically significant positive association
with PFBS exposure. No studies have been identified that evaluate the association between
PFBS exposure and potential cancer outcomes. While the epidemiology studies were not
influential to drawing evidence integration judgments or the derivation of toxicity values, the
general findings identify potential areas of future research.
Animal studies of repeated-dose PFBS exposure have been exclusively via the oral route,
used the potassium salt of PFBS (K+PFBS) as the source exposure material, and have examined
noncancer effects only. The available rat and mouse studies support identification of thyroid,
developmental, and kidney endpoints as potential health effects following repeated exposures in
utero and/or during adulthood. Animal studies have also evaluated other health outcomes, such
as liver effects, reproductive parameters, lipid/lipoprotein homeostasis, and effects on the spleen
and hematology; however, the available evidence does not support a clear association with PFBS
exposure and these outcomes.
Noncancer Effects Observed Following Oral Exposure
Oral exposures to PFBS or its K+ salt in adult and developing rats and mice have been
shown to result in thyroid, developmental, and kidney effects. Thyroid effects in exposed adult
rats and mice and in developing mice were primarily expressed through significant decreases in
circulating levels of hormones such as thyroxine (T4) and triiodothyronine (T3). In early
developmental life stages in mice (e.g., newborn), decreases in thyroid hormone were
accompanied by other effects indicative of delayed maturation or reproductive development
(e.g., vaginal patency and eyes opening). Kidney weight and/or histopathological alterations
(e.g., renal tubular and ductal epithelial hyperplasia) were observed in rats following short-term
and subchronic oral exposures. Many of the kidney effects, however, occurred at higher doses
than did the thyroid and developmental effects. The limited number of human studies examining
oral PFBS exposure does not inform the potential for effects in thyroid, developing offspring, or
the renal system.
Oral Reference Doses for Noncancer Effects
Subchronic1 and chronic2 oral RfDs were derived for PFBS. The hazards of potential
concern include thyroid, developmental, and kidney effects. From these identified targets of
PFBS toxicity, perturbation of thyroid hormone levels (e.g., T4) was used as the critical effect for
deriving a subchronic and chronic RfD. Based on recommendations in the U.S. EPA's
Recommended Use of Body Weight3/4 as the Default Method in Derivation of the Oral Reference
Dose (U.S. EPA. 2011b). chemical-specific toxicokinetic data (e.g., serum half-lives) were used
to scale a toxicologically equivalent dose of orally administered PFBS from animals to humans.
Following the U.S. EPA's Benchmark Dose Technical Guidance Document (U.S. EPA. 2012).
benchmark dose (BMD) modeling of thyroid effects in a developmental life stage following
exposure to K+PFBS in utero resulted in a BMDLo.ssd human equivalent dose (FLED) of
Subchronic exposure: Repeated exposure by the oral, dermal, or inhalation route for more than 30 days, up to
approximately 10% of the lifespan in humans (more than 30 days up to approximately 90 days in typically used
laboratory animal species).
2Chronic exposure: Repeated exposure by the oral, dermal, or inhalation route for more than approximately 10% of
the lifespan in humans (more than approximately 90 days to 2 years in typically used laboratory animal species).
(https://ofmpub.epa.gov/sor internet/registrv/termreg/searchandretrieve/glossariesandkevwordlists/search.do?details
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EPA/690/R-21/001F
0.095 milligrams per kilogram per day (mg/kg-day). This HED associated with thyroid effects
served as the point of departure (POD) for deriving the subchronic and chronic RfDs.
The subchronic RfD for K+PFBS was calculated by dividing the POD (HED) for
decreased serum total T4 observed in newborn (Postnatal Day [PND] 1) mice, in the study
conducted by Feng et al. (20171 by a composite uncertainty factor (UFc) of 100 to account for
extrapolation from mice to humans (an interspecies uncertainty factor, or UFa, of 3), for
interindividual differences in human susceptibility (intraspecies uncertainty factor, or UFh, of
10), and for deficiencies in the toxicity database (database uncertainty factor, or UFd, of 3) (a
value of 1 was applied for subchronic-to-chronic UF, or UFs, and LOAEL-to-NOAEL
uncertainty factor, or UFl) (see Table 10), yielding a subchronic RfD of 0.00095 mg/kg-day
rounded to 1 x 10 3 mg/kg-day. Because K+PFBS is fully dissociated in water at the
environmental pH range of 4-9 to the PFBS anion (PFBS ) and the K+ cation, data for K+PFBS
were used to derive a subchronic RfD for the free acid (PFBS) by adjusting for differences in
molecular weight (MW) between K+PFBS (338.19) and PFBS (300.10), yielding the value of
0.00085 mg/kg-day rounded to 9 x 10 4 mg/kg-day for the subchronic RfD for PFBS (free acid).
The chronic RfD for K+PFBS associated with thyroid effects was calculated by dividing
the POD (HED) for decreased serum total T4 observed in newborn (PND 1) mice, in the study
conducted by Feng et al. (20171 by a UFc of 300 to account for extrapolation from mice to
humans (UFa of 3), for interindividual differences in human susceptibility (UFh of 10), and
deficiencies in the toxicity database (UFd of 10) (a value of 1 was applied for UFs and UFl) (see
Table 12), yielding a chronic RfD of 0.00032 mg/kg-day rounded to 3 x 10 4 mg/kg-day. Like
the subchronic RfD for thyroid effect, based on the data for K+PFBS, a chronic RfD for PFBS
(free acid) of 0.00028 mg/kg-day rounded to 3 x 10 4 mg/kg-day was derived.
Confidence in the Oral RfDs
The overall confidence in the subchronic RfD for thyroid effects is medium. The
gestational exposure study conducted by Feng et al. (2017) reported administration of K PFBS
by gavage in pregnant Institute of Cancer Research (ICR) mice (10/dose) from Gestation Days
(GDs) 1 to 20. This study was of good quality (i.e., high confidence) with adequate reporting
and consideration of appropriate study design, methods, and conduct (click to see risk of bias
analysis in HAWC3). Confidence in the oral toxicity database for derivation of the subchronic
RfD is medium because, although there are multiple short-term studies and a
sub chronic-duration toxicity study in laboratory animals, a two-generation reproductive toxicity
study in rats (l.ieder et al.. 2009b). and multiple developmental toxicity studies in mice and rats,
there are no PFBS studies available that have specifically evaluated health effect domains of
emerging concern across the PFAS class such as immunotoxicity and mammary gland
development (Dewitt et al.. 2012; White et al.. 2007). Further, neurodevelopmental effects are of
particular concern when perturbations in thyroid hormone occur during a sensitive early life
stage, and the absence of a study evaluating neurodevelopmental effects following PFBS
exposure is a source of uncertainty in the assessment.
The overall confidence in the chronic RfD for thyroid effects is low. Although the
chronic RfD, like the subchronic RfD, was derived using data from the high-confidence principal
study conducted by Feng et al. (2017). there is increased concern about the potential for
3HAWC: A modular web-based interface to facilitate development of human health assessments of chemicals; see
Appendix D for details.
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EPA/690/R-21/001F
identification of hazards following longer (i.e., chronic) duration PFBS exposures. In addition,
because of the lack of studies that specifically evaluated health effect domains of emerging
concern across the PFAS class, such as immunotoxicity, mammary gland development, or
neurodevelopmental at any exposure duration—but particularly for chronic duration—
confidence in the database specifically for a chronic RfD is low.
Effects Other Than Cancer Observed Following Inhalation Exposure
There are no studies available that examined toxicity in humans or experimental animals
following inhalation exposure, thereby precluding the derivation of an inhalation reference
concentration (RfC).
Evidence for Carcinogenicity
Under the U.S. EPA's Guidelines for Carcinogen Risk Assessment (U.S. EPA. 2005). the
Agency concluded that there is "Inadequate Information to Assess Carcinogenic Potential" for
PFBS and K+PFBS by either oral or inhalation routes of exposure. Therefore, the lack of data on
the carcinogenicity of PFBS and the related compound K+PFBS precludes the derivation of
quantitative estimates for either oral (oral slope factor) or inhalation (inhalation unit risk)
exposure.
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EPA 690 R-21 00IF
1.0 BACKGROUND
1.1 PHYSICAL AND CHEMICAL PROPERTIES
Perfluorobutane sulfonic acid (PFBS) (Chemical Abstracts Service registry number
[CASRN] 375-73-5)4 and its related salt, potassium perfluorobutane sulfonate (K+PFBS)
(CASRN 29420-49-3), are members of the group of per- and polyfluoroalkyl substances (PFAS),
more specifically the short-chain perfluoroalkane sulfonates. For purposes of this assessment,
"PFBS" will signify the ion, acid, or any salt of PFBS. Concerns about PFBS and other PFAS
stem from the resistance of these compounds to hydrolysis, photolysis, and biodegradation,
which leads to their persistence in the environment (Sundstrom et al., 2012). The chemical
formula of PFBS is C4HF9O3S and the chemical formula of K+PFBS is C4F9KO3S. Their
respective chemical structures are presented in Figure 1. K+PFBS differs from PFBS by being
associated with a potassium ion. The reported water solubility of each species suggests that in
aqueous environments the sulfonate would be the predominant form. The preferential use of
K+PFBS in laboratory studies is related to the optimal dissociation of the salt to the sulfonate
(i.e., PFBS") at pH values ranging from 4 to 9 (see Table 1). Table 1 provides a list of the
physicochemical properties for PFBS and K+PFBS.
K*
PFBS
K+PFBS
Figure 1. PFBS and K+PFBS Chemical Structures
4The CASRN given is for linear PFBS; the source PFBS used in toxicity studies was assayed at >98% linear,
suggesting some minor proportion of other chemicals, such as branched PFBS isomers, are present. Thus, observed
health effects may apply to the total linear and branched isomers in a given exposure source.
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EPA/690/R-21/001F
Table 1. Physicochemical Properties of PFBS (CASRN 375-73-5) and Related Compound
K+PFBS (CASRN 29420-49-3)
Property (unit)
Value3
PFBS (free acid)b
K+PFBS (potassium salt)0
Boiling point (°C)
152
447
Density (g/cm3)
1.83 (predicted)
1.83 (predicted)
Vapor pressure (mm Hg)
0.104 (predicted)
1.12 x 10-8
pH
ND
ND
Solubility in water (mol/L)
0.0017
0.08
Molecular weight (g/mol)
300.09
338.18
Dissociation constant
NA
Fully dissociated in water over the pH range of 4-9
aValues are experimentally determined unless otherwise indicated.
bU.S. EPA Chemistry Dashboard for CASRN 375-73-5.
°U.S. EPA Chemistry Dashboard for CASRN 29420-49-3.
K+PFBS = potassium perfluorobutane sulfonate; NA = not applicable; ND = no data; PFBS = perfluorobutane
sulfonic acid.
1.2 OCCURRENCE
PFBS-based compounds are surfactants used primarily in the manufacture of paints,
cleaning agents, and water- and stain-repellent products and coatings. They serve as
replacements for periluorooctane sulfonic acid (PFOS) (3M, 2002b). Various sources report
detection or occurrence in environmental media and consumer products, including drinking
water, ambient water, dust, carpeting and carpet cleaners, floor wax, and food packaging. To
assess potential health risks associated with these occurrences, an exposure assessment, which is
beyond the scope of this document, would be necessary to determine the relative source
contribution to human PFBS exposure from each reported occurrence and the relevance, if any,
to human health.
Oral exposure via drinking water might be expected in areas where contamination has
been reported. U.S. EPA Unregulated Contaminant Monitoring Rule data for public drinking
water utilities in 2013-2015 showed levels of PFBS above the minimum reporting level
(>0.09 micrograms per liter [jug/L]) in water systems serving Alabama, Colorado, Georgia, the
Northern Mariana Islands, and Pennsylvania (U.S. HP A. 2017; I hi et al.. 2016). These utilities
used both ground and surface drinking water sources, with PFBS concentrations ranging from
0.09 to 0.37 |ig/L. The estimated combined number of people served by these water systems is
more than 340,000 (U.S. EPA. 2018).
Measurements from 37 surface water bodies in the northeastern United States
(metropolitan New York area and Rhode Island) collected in 2014 showed an 85% site detection
rate (Zhang et al.. 2016). PFBS has also been identified in surface waters in Georgia, New
Jersey, North Carolina, and the Upper Mississippi River Basin (Post et al.. 2013; Easier et al..
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EPA/690/R-21/001F
2011; Nakavama et al.. 2010; Nakavama et al.. 2007). It has also been detected in wastewater
treatment plant effluent, seawater, soil, and biosolids (Houtz et al.. 2016; Zhao et al.. 2012;
Sepulvado et al.. 2011).
PFBS contamination, which has been associated with the use of aqueous film-forming
foams (AFFFs) (ESTCP. 2017; Anderson et al.. 2016). was reported at Superfund sites and areas
under assessment for Superfund designation. Contaminated sites include the former Wurtsmith
Air Force Base, Ellsworth Air Force Base, and Dover Air Force Base (Acrostar SES LLC. 2017;
Anonymous. 2017; ASTSWMO. 2015). At the Wurtsmith site, PFBS was detected at a
concentration of 6.4 |ig/L in groundwater contaminated by a PFAS plume originating from the fire
training area (ASTSWMO. 2015). It is also present in some drinking water samples from nearby
residential wells at low nanograms per liter concentrations, which were below the screening value
cited by the Michigan Department of Community Health (MDCH. 2015). Other sources of PFAS
and/or PFBS contamination include chrome plating operations, PFAS manufacture, and sites that
use PFAS in product formulations such as textile and electronics facilities (Wang et al.. 2013).
PFBS has also been detected in household dust and consumer products. There was a 92%
detection frequency for PFBS among 39 household dust samples (10 from the United States)
analyzed with levels ranging from 86 nanograms per gram (ng/g) for the 25th percentile to
782 ng/g for the 75th percentile (Kato et al.. 2009). In a separate study, PFBS dust levels were
measured in Boston area offices (n = 31), homes (n = 30), and vehicles (n = 13) with detection
frequencies being relatively low—10, 3, and 0%, respectively—and ranging in the low parts per
billion (Fraser et al.. 2013). Consumer products could also be an exposure source. Limited
quantitative testing showed the presence of PFBS in carpet and upholstery protectors (45.8 and
89.6 ng/g), carpet shampoo (25.7 and 911 ng/g), textiles (2 ng/g), and floor wax (143 ng/g)
purchased in the United States (Liu et al.. 2014).
PFBS is not authorized for use in food packaging. However, PFBS was detected in fast
food packaging (7/20 samples) in one U.S. study (Sellaider et al.. 2017) although the magnitude
of the detection was not reported.
The European Food Safety Authority reported the presence of PFBS in various food and
drink items, including fruits, vegetables, cheese, and bottled water. For average adult
consumers, the estimated exposure ranges for PFBS were 0.03-1.89 nanograms per kilogram per
day (ng/kg-day) (minimum) to 0.10-3.72 ng/kg-day (maximum) (EFSA. 2012).
PFBS has been reported in serum of humans in the general population. In American Red
Cross samples collected in 2015, 8.4% had a quantifiable serum PFBS concentration; the
majority of samples were below the lower limit of quantitation (4.2 nanograms per milliliter
[ng/mL]) (01 sen et al.. 2017). The National Health and Nutrition Examination Survey
(NHANES) included PFBS in consecutive biomonitoring cycles, including 2013-2014 where the
95th percentile reported for PFBS was at or below the level of detection (0.1 ng/mL).
Considering the relatively rapid rate of elimination of PFBS (days to weeks), compared with
longer chain PFAS (years), the lack of biomonitoring detects (e.g., NHANES 2013-2014 cycle)
should not be interpreted as a lack of occurrence or exposure potential. Another study with a
lower limit of detection (0.013 ng/g) reported increasing levels of PFBS in serum from
pri mi parous nursing women in Sweden from 1996 to 2010 (Glvnn et al.. 2012).
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1.3 TOXICOKINETICS
1.3.1 Overview
Animal evidence has shown that PFBS, like other PFAS, is well absorbed following oral
administration. PFBS distributes to all tissues of the body (Bogdanska et al.. 2014). but a study
evaluating the volume of distribution (Fd) concluded that distribution is predominantly
extracellular (01 sen et al .. 2009). Because of its resistance to metabolic degradation, PFBS is
primarily eliminated unchanged in urine and feces.
Three sets of investigators have conducted toxicokinetic studies in rats and monkeys
(Huang et al.. 2019a; Chengelis et al.. 2009; 01 sen et al.. 2009). 01sen et al. (2009) and Xu et al.
(2020) have measured the half-life of PFBS in humans. Bogdanska et al. (2014) and Lau et al.
(2020) have reported limited toxicokinetic information in mice. One study developed a
physiologically based pharmacokinetic (PBPK) model that includes parameterization for PFBS
(Fabrega et al.. 2015).
Results of all studies discussed in this section are summarized in Table 2.
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EPA/690/R-21/001F
Table 2. Summary of the Toxicokinetics of Serum PFBS (Mean ± SE)
Species/Sex
Study Design
Elimination
Half-Life (hr)
AUC
(jig-hr/mL)
Clearance
Vd
(L/kg)
Reference
Mice
Mice/male
Single oral dose (30 mg/kg)
3.7
1,515
0.019 (L/hr-kg)
0.129
Lau et al. (2020)
Single oral dose (300 mg/kg)
6.0
7,178
0.039 (L/hr-kg)
0.291
Lau et al. (2020)
Single oral dose
(combined 30/300 mg/kg)
5.8

0.038 (L/hr-kg)
0.275
Lau et al. (2020)
Mice/female
Single oral dose (30 mg/kg)
4.4
520
0.056 (L/hr-kg)
0.145
Lau et al. (2020)
Single oral dose (300 mg/kg)
4.6
4,587
0.064 (L/hr-kg)
0.308
Lau et al. (2020)
Single oral dose
(combined 30/300 mg/kg)
4.5

0.063 (L/hr-kg)
0.278
Lau et al. (2020)
Rats
Rats/male
Single i.v. dose (10 mg/kg)
2.1
254
0.0394 (L/hr-kg)
0.118
Chengelis et al. (2009)
Single i.v. dose (30 mg/kg)
4.51 ± 2.22a
294 ± 77
119 ± 34 (mL/hr)b
0.330 ±0.032
Olsen et al. (2009)
Single oral dose (30 mg/kg)
4.68 ± 0.43a
163 ± 10
NA
0.676 ±0.055
Olsen et al. (2009)
Single i.v. dose (4 mg/kg)
4.22 ± 0.28°
116 ± 7
0.0345 ± 0.002 (L/hr-kg)
0.188 ±0.017c
Huang et al. (2019a)
Single oral dose (4 mg/kg)
4.89 ± 1.67°
154 ± 15
0.0265 ± 0.003 (L/hr-kg)
0.174 ±0.614c
Huang et al. (2019a)
Single oral dose (20 mg/kg)
5.36 ± 1.24°
533 ±45
0.0376 ± 0.003 (L/hr-kg)
0.167 ±0.039c
Huang et al. (2019a)
Single oral dose (100 mg/kg)
5.25 ± 1.19°
1,320 ± 100
0.0755 ± 0.006 (L/hr-kg)
0.335 ±0.041c
Huang et al. (2019a)
Rats/female
Single i.v. dose (10 mg/kg)
0.64
32
0.311 (L/hr-kg)
0.288
Cheneelis et al. (2009)
Single i.v. dose (30 mg/kg)
3.96 ± 0.21a
65 ±5
469 ± 40 (mL/hr)d
0.351 ±0.034
Olsen et al. (2009)
Single oral dose (30 mg/kg)
7.42 ± 0.79a
85 ± 12
NA
0.391 ±0.105
Olsen et al. (2009)
Single i.v. dose (4 mg/kg)
0.95 ± 0.10°
16 ± 1
0.252 ± 0.018 (L/hr-kg)
0.165 ±0.015c
Huang et al. (2019a)
Single oral dose (4 mg/kg)
1.50 ± 0.10°
29 ±3
0.152 ±0.020 (L/hr-kg)
0.328 ±0.042c
Huang et al. (2019a)
Single oral dose (20 mg/kg)
1.23 ± 0.12°
109 ± 23
0.183 ±0.039 (L/hr-kg)
0.326 ±0.073c
Huang et al. (2019a)
Single oral dose (100 mg/kg)
l.ll±0.10c
387 ± 50
0.259 ± 0.033 (L/hr-kg)
0.415 ±0.063c
Huang et al. (2019a)
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Table 2. Summary of the Toxicokinetics of Serum PFBS (Mean ± SE)
Species/Sex
Study Design
Elimination
Half-Life (hr)
AUC
(jig-hr/mL)
Clearance
Vd
(L/kg)
Reference
Monkeysb
Cynomolgus
macaque/male
Single i.v. dose (10 mg/kg)
15 (9.65)e
1,115 ±859
0.016 (L/hr-kg)
0.209 ± 0.028
Chengelis et al. (2009)
Single i.v. dose (10 mg/kg)
95.2 ±27.1
24.3 ± 8.6
511 ± 141 (mL/hr)
0.254 ±0.031
Olsen et al. (2009)
Cynomolgus
macaque/female
Single i.v. dose (10 mg/kg)
8.1
489±180
0.0229 ± 0.0099 (L/hr-kg)
0.248 ± 0.045
Chengelis et al. (2009)
Single i.v. dose (10 mg/kg)
83.2 ±41.9
35.4 ± 13.3
368 ± 120 (mL/hr)
0.255 ±0.017
Olsen et al. (2009)
Humans
Males and female
Occupational (n = 6)
619.2f
NA
NA
NA
Olsen et al. (2009)
Males
Occupational (n =5)
552f
NA
NA
NA
Olsen et al. (2009)
Female
Occupational (n = 1)
1,096.8
NA
NA
NA
Olsen et al. (2009)
Males and
females
Occupational (n = 26)
1,056
NA
NA
NA
Xu et al. (2020)
a01sen et al. (2009) reported t\n,a and /i/2,p in rats, presenting data for tw.fi.
bBody weights were reported to be 0.200-0.250 kg (with corresponding clearance of approximately 476 mL/hr-kg).
°Huang et al. (2019a) reported tn2,a, /i/2,p, and /i/2kio in male rats (both oral and i.v.) and female rats (i.v. only); only /lokio was reported in female rats (oral).
Presenting data for h 2.|; for male rats (both oral and i.v.) and female rats (i.v.) and /12km for female rats (oral). The volume of distribution (I '1) was calculated
as the sum of volume terms of the central compartment and that of the peripheral compartment except for orally exposed female rats. The volume of the
peripheral compartment was not reported for orally exposed female rats, representing the volume of the central compartment only.
'The data were monitored 48 hours and 31 days postdosing for Chengelis et al. (2009) and Olsen et al. (2009). respectively.
eOne male monkey had a serum concentration more than 10-fold higher than the others at 48 hours postdosing with an estimated PFBS half-life of 26 hours.
'Olsen et al. (2009) reported mean and geometric mean values for males only and all subjects, presenting data for geometric mean values.
AUC = area under the curve; i.v. = intravenous; NA = not available; PFBS = perfluorobutane sulfonic acid; SE = standard error; ty2 = half-life; Vd = volume of
distribution.
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1.3.2 Absorption
01 sen et al. (2009) conducted intravenous (i.v.) and oral uptake studies in rats (n = 3/sex)
that were given a single dose (30 milligrams per kilogram [mg/kg]) of potassium PFBS
(K+PFBS). The serum area under the concentration curve (AUC) after i.v. administration was
294 ± 77 and 65 ± 5 (|ig-hour/mL) in male and female rats, respectively, and 163 ± 10 and
85 ± 12 in males and females, respectively, after oral dosing. The large variance in AUC for
male rats after i.v. dosing and greater AUC after oral dosing compared to i.v. dosing in females
makes it difficult to interpret these results with certainty, but it seems that PFBS is 100%
bioavailable in female rats, whereas the nominal bioavailability in male rats is only 55% based
on AUC. Peak concentrations (Cmax) occurred at 0.3-0.4 hours after oral dosing, showing that
absorption was fairly rapid. Bioavailability based on Cmax was 60% in male rats and 85% in
female rats, suggesting a similar sex difference as estimated from the AUCs.
The above findings are generally confirmed by Huang et al. (2019a) who found that
absorption of PFBS usually occurred within 24 hours, along with the time reaching the maximal
plasma concentration (Jmax) under 2.4 hours in male rats and under 1.4 hours in female rats,
following a single dose of gavage administration in Hsd:Sprague-Dawley (S-D) rats (4, 20,
100 mg/kg of K+PFBS). However, bioavailability calculated based on the AUC after i.v. and
oral doses of 4 mg/kg reported by Huang et al. (2019a) was 75% in males and 60% in females.
The Cmax values of 45% and 27% in males and females, respectively, are qualitatively the
opposite of the results from 01 sen et al. (2009).
Given the range of estimated bioavailability from the results of 01 sen et al. (2009) and
Huang et al. (2019a). a difference in this parameter between male and female rats cannot be
determined. Averaging the AUC-based values for both males and females from the two studies
yields an overall average of 73%.
Notably, Huang et al. (2019a) also observed that the dose-adjusted AUC decreased with
increasing doses for both males and females. However, this result could be attributed to
saturation of renal resorption at higher doses, rather than a reduction in absorption.
Similar observations indicating rapid absorption of PFBS have been reported for CD-I
mice orally exposed to PFBS at 30 or 300 mg/kg, where '/max was estimated to occur between
1 and 2 hours after gavage (Lau et al.. 2020).
1.3.3 Distribution
PFBS has been shown to distribute to tissues within 24 hours of exposure, with the liver
and kidney being the organs with highest distribution. Lau et al. (2020) evaluated the
pharmacokinetic properties of PFBS in CD-I mice at 8 weeks of age. Male and female mice
were given a single dose of 0, 30, or 300 mg/kg body weight PFBS via gavage. The liver and
kidneys were harvested 24 hours postdosing. PFBS distributed to both organs readily in a
dose-dependent manner but did not accumulate in either organ. Lau et al. (2020) reported
similar combined Vd values of 0.275 or 0.278 liter per kilogram [L/kg] in male and female mice,
respectively (Table 2).
01 sen et al. (2009) estimated volumes of distribution for K PFBS as 0.7 and 0.4 L/kg in
male and female rats, respectively, and 0.25 L/kg in male and female cynomolgus macaques and
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concluded that K+PFBS is primarily distributed in the extracellular space. Consistent with the
observations by 01 sen et al. (20091 Huang et al. (2019a) found that the overall Vd for PFBS was
generally comparable between male rats (0.167-0.335 L/kg) and female rats (0.165-0.415 L/kg).
Chengelis et al. (2009) calculated a Vd of 0.248 L/kg in female cynomolgus macaques, consistent
with females from 01 sen et al. (2009). The male monkey Vd from Chengelis et al. (2009) was
slightly lower (0.209 L/kg) than corresponding females and males from 01 sen et al. (2009).
These results indicate Vd is generally comparable between male and female monkeys. Huang et
al. (2019a) also evaluated tissue concentrations in the liver, kidney, and brain of male and female
rats and reported higher PFBS concentrations in the liver than the kidney, with the lowest
concentrations occurring in the brain.
Bogdanska et al. (2014) characterized the tissue distribution of 35S-labeled PFBS in male
C57BL/6 mice. The animals (3/group) were exposed for either 1, 3, or 5 days to an average of
16 mg of PFBS/kg-day in the diet. Following 1, 3, and 5 days of exposure, the total estimated
recovery of PFBS from all tissues evaluated was 10, 5, and 3.4% of the ingested dose,
respectively. The declining recovery with time reflects the lack of accumulation in tissues after
the first few days, with continued elimination in the urine. The study authors suggested that
these low recovery rates most likely reflect rapid excretion of PFBS and/or potentially limited
uptake of the compound; however, the results of I .au et al. (2020) and 01 sen et al. (2009) suggest
that limited tissue distribution is also a factor.
Bogdanska et al. (2014) found that blood levels of PFBS did not change when comparing
values observed after 1 and 5 days of exposure. As with PFOS, PFBS was found to distribute to
most of the 20 tissues examined at all exposure durations, but the levels of PFBS were
significantly lower (fivefold to 40-fold lower) than those of PFOS in tissues after similar
exposure to PFOS, especially in liver and lungs (Bogdanska et al.. 2014). These differences
might be attributed to chain-length-dependent active transport of perfluorinated chemicals
(Weaver et al.. 2010). Excluding stomach and fat tissue, PFBS tissue levels increased between
1 and 3 days of exposure, but there were no significant changes in tissue levels between 3 and
5 days of exposure in any tissue examined. As with PFOS, whole bone, liver, blood, skin, and
muscle accounted for approximately 90% of the recovered PFBS at all time points. The highest
tissue concentrations outside of blood, however, were found in the liver, GI tissues, kidney, and
cartilage. The significant total PFBS mass found in muscle and skin was due to the large total
volume of these tissues rather than the per unit concentration in them. The liver contained the
highest tissue concentration of PFBS at all time points, while the brain contained the lowest.
Human studies were not available on lactational transfer of PFBS. Studies are sparse
pertaining to the transplacental transfer of PFBS in humans; in a Spanish mother-child paired
cohort, PFBS was not found in maternal blood samples or in corresponding cord blood during
the first trimester of pregnancy (Manzano-Salgado et al .. 2015). However, developmental
studies in animals indicate the potential for effects in offspring following gestational exposure,
suggesting direct (i.e., fetus) and/or indirect (maternal/pregnant dam) effects of PFBS on
offspring (Feng et al.. 2017; York. 2003a. 2002).
Volume of distribution is expected to be similar across mammalian species. For PFBS,
the average value for male and female monkeys (0.23 L/kg) is in the range estimated for male
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and female rats by Huang et al. (2019a) (0.17-0.42 L/kg), although estimates by 01 sen et al.
(2009) were slightly higher.
1.3.4 Metabolism
There is no evidence of biotransformation of PFBS. It is expected that PFBS, a
short-chain (C4) of perfluoroalkyl acids (PFAAs), is metabolically inert because of the chemical
stability that also exists in the longer chain PFAA chemicals, including perfluorohexane sulfonic
acid (PFHxS) (C6), PFOS (C8), and perfluorooctanoic acid (PFOA) (C8).
1.3.5 Excretion
To facilitate comparison of differing studies for a given species, results for excretion are
organized by species.
1.3.5.1 Mice
Lau et al. (2020) dosed male and female CD-1 mice with 0, 30, or 300 mg/kg body
weight PFBS via a single gavage dose. Trunk blood was collected at 0.5, 1, 2, 4, 8, 16, 24, and
48 hours after dosing and urine at 24 hours after dosing. Within 24 hours of dosing, more than
95% of the PFBS measured in serum was excreted into urine. Although the rate of PFBS
clearance was linear with administered doses, urine accounted for only 30-43% of the original
gavage doses. The half-life of PFBS was estimated to be 4.5 hours in the female mice and
5.8 hours in the males. Sex difference in PFBS elimination was also noted in that the elimination
rate of absorbed PFBS was about 28% faster in female mice than male mice. Similarly, AUC
estimates for the serum, kidney, and liver compartments were higher in males than in females.
The findings are generally comparable to previous studies on rats (Huang et al.. 2019a; 01 sen et
al. 2009).
1.3.5.2 Rats
Chengelis et al. (2009) conducted a single-dose pharmacokinetic study in S-D rats,
designed to compare the toxicokinetic behavior of PFBS with that of perfluorohexanoic acid
(PFHxA), another PFAA. In this study, 12 male and 12 female rats were each administered a
bolus dose of PFBS (10 mg/kg) via i.v. injection. Blood samples were collected from
three animals per sex at 0.5, 1, 1.5, 2, 4, 8, and 24 hours after dosing. Additionally, to determine
urinary excretion, three animals per sex were housed in metabolic cages following dose
administration and their urine collected over the following time intervals: 0-6, 6-12, and
12—24 hours postdosing. Chengelis et al. (2009) fit the data to a noncompartmental model to
calculate pharmacokinetic parameters. Female rats had an approximately threefold shorter mean
elimination half-life of PFBS in serum (0.64 hour) than male rats (2.1 hour). This result could be
in part due to the difference in clearance and Vd. The mean apparent clearance of PFBS from the
serum was approximately eightfold higher for female rats (0.311 L/hour-kg) than for male rats
(0.0394 L/hour-kg), and the mean apparent Vd for PFBS in the serum was approximately 2.4-fold
higher for female rats (0.288 L/kg) than for male rats (0.118 L/kg). Approximately 70% of the
administered dose of PFBS was recovered in the urine over 24 hours postdosing regardless of
sex. Using the urine data, the mean half-life values for male rats and female rats were
determined to be 3.1 and 2.4 hours, respectively; the finding of longer urinary half-lives in males
is consistent with those observed for serum half-lives.
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01 sen et al. (2009) evaluated the elimination of PFBS in S-D rats after i.v. and oral
exposure to K+PFBS. The terminal serum elimination half-lives following i.v. administration of
30 mg/kg K+PFBS were 4.51 ± 2.22 hours for males and 3.96 ± 0.21 hours for females
(mean ± standard deviation [SD]). Although there was no statistically significant difference
between the terminal serum half-lives in male and female rats, there was a statistically significant
difference in the urinary clearance rates (p < 0.01), with female rats (469 ± 40 mL/hour) having
faster clearance rates than male rats (119 ± 34 mL/hour). Because clearance [CL] is calculated
from the ratio of the volume of distribution [Vd] to the half-life [tv 2], CL = 0.693 x Vd ^ tin,
differences in Vd can lead to differences in CL, even when l\ 2 is similar between comparison
groups. For rats receiving an oral dose, terminal serum K+PFBS elimination half-lives were
significantly different (p < 0.05) for males (tv2 = 4.68 ± 0.43 hour) versus females
(tv2 = 7.42 ± 0.79 hour).
Huang et al. (2019a) also evaluated elimination of PFBS following a single i.v. or gavage
dose in male or female Hsd:S-D rats (4, 20, 100 mg/kg of K+PFBS). They reported elimination
half-lives (tv 2,p) following i.v. administration of PFBS in male and female rats of 4.22 and
0.95 hours, respectively. The data for male rats after both oral and i.v. dosing and female rats
administered PFBS by i.v. fit a two-compartment model, whereas data in female rats dosed via
gavage fit a one-compartment model. Thus, elimination half-lives were only reported for male
rats following oral exposure and ranged from 4.89-5.36 hours. Overall plasma elimination
half-lives (tin kio) reported in female rats after oral administration were between
1.11-1.50 hours, approximately three to fourfold faster than in males that ranged from
4.89-5.36 hours. Similarly, clearance was three to sixfold higher in females than males given
the same dose (26.5-75.5 mL/hour-kg in males, 152-259 mL/hour-kg in females).
The serum K PFBS elimination half-lives reported by Huang et al. (2019a) are consistent
with the findings of 01 sen et al. (2009) in male rats but not in female rats. In general, the
elimination half-life of serum PFBS observed by Huang et al. (2019a) in female rats was two to
fourfold shorter than seen by 01 sen et al. (2009). Similarly, Chengelis et al. (2009) calculated
half-lives using a one compartment model for each group, whereas 01 sen et al. (2009)
determined separate a and P phases via a two-compartment model. Thus, the half-life estimates
of 01 sen et al. (2009) following i.v. administration (4.5 1-3.96 hours) are higher than those
estimated by Chengelis et al. (2009) based on urine data (0.64-2.1 hours).
1.3.5.3 Monkeys
Similar to their study in rats, Chengelis et al. (2009) investigated the toxicokinetic profile
of PFBS through a series of experiments in the cynomolgus macaque (Macaca fascicularis).
Monkeys (three males and three females) were each administered a bolus i.v. dose of 10 mg/kg
PFBS. The controlled exposure to PFBS occurred 7 days after the same animals were each
administered a bolus dose of PFHxA (10 mg/kg). Blood samples were collected at 0 hours
(immediately prior to dosing) and at 1, 2, 4, 8, 24, and 48 hours after dose administration and
were analyzed to determine PFBS concentration in serum. Only a single clearance half-life was
estimated. The estimated half-life of PFBS in serum ranged from 5.8 to 26.0 hours in this
experiment, and the median half-life was 9.55 hours for the six animals.
01 sen et al. (2009) also evaluated the elimination of PFBS (specifically, K PFBS) in
cynomolgus macaques after i.v. dosing. A significant difference in design from the study of
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Chengelis et al. (2009) is that 01 sen et al. (2009) followed PFBS elimination for 31 days in
monkeys (vs. 48 hours), allowing them to identify both an initial clearance half-life and a
terminal-phase half-life. 01 sen et al. (2009) did not observe statistically significant sex-related
differences in half-life or clearance between male and female monkeys, unlike those observed in
rats. In monkeys, the mean terminal serum elimination half-lives, after i.v. administration of
10 mg/kg K+PFBS, were 95.2 ± 27.1 hours in males and 83.2 ± 41.9 hours in females.
The serum half-life data in 01 sen et al. (2009) clearly show a slow elimination phase in
monkeys that does not begin until 4-10 days after dosing. Chengelis et al. (2009) followed
elimination for only 48 hours, hence could not have observed this terminal clearance phase. The
initial elimination half-life (7i/2,p) estimated by 01 sen et al. (2009) in monkeys—13 hours for
males, 11 hours for females—is essentially identical to the values estimated by Chengelis et al.
(2009)—10 or 15 hours for males (without/with outlier) and 8 hours in females. Hence the two
studies appear consistent in identifying an initial elimination half-life, but the difference in
design precluded Chengelis and colleagues from identifying the longer (terminal) half-life of
PFBS.
1.3.5.4 Humans
In addition to their experimental studies in rats and monkeys, 01 sen et al. (2009)
evaluated the elimination of human serum K+PFBS in a group of workers with occupational
exposure, with serum concentrations measured up to 180 days after cessation of further K+PFBS
work-related activity. Given that the workers had been occupationally exposed, distribution into
the tissues is expected to have been complete before the observations began. The reported mean
serum half-life was 23 days in males (n = 5) and 45.7 days in females (n = 1). Among the six
subjects (five males, one female), the reported geometric mean serum elimination half-life for
K+PFBS was 25.8 days (95% confidence interval [CI]: 16.6-40.2 days). Because there was only
one female subject, these data cannot be used to establish a significant sex difference in
elimination. Urine appeared to be a major route of elimination in humans based on observed
urine levels of PFBS in the study.
Xu et al. (2020) also measured PFBS elimination in a study population with previous
occupational exposure, in this case airport employees who were exposed to firefighting foam that
contained PFBS. Eleven male and six female employees provided repeated blood samples
during a period of observation with minimal exposure, and the data were analyzed with a linear
mixed-effects pharmacokinetic model. The average half-life was 44 days (95% CI: 37-55 days).
Although Xu et al. (2020) evaluated age and sex as covariates of their statistical model, they did
not report either as being a significant factor for PFBS elimination. The average half-life
(44 days) is larger than that reported by 01 sen et al. (2009) (25.8 days), but there is significant
overlap: the range of Xu et al. (2020) is 21.6-87.2 days while the range of 01 sen et al. (2009) is
13.1-45.7 days.
For the sake of comparison, the linear mixed model used by Xu et al. (2020) was also
applied to the estimated serum PFBS elimination half-life for the population and each individual
worker (five male, one female) who manufactured K PFBS, described in 01 sen et al. (2009). In
brief, a linear mixed effect model is an extension of simple linear models that can be used to
estimate toxicokinetic parameters such as the serum elimination rate constant (keiim) and half-life
by assuming one-compartment first-order elimination kinetics. The details of the linear
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mixed-effect model have been reported previously Li et al. (2018). Because of the limited
sample size (only one female worker) and the lack of data on participant age for each worker in
the study, age and sex were not included in the linear mixed model for reanalysis of the 01 sen et
al. (2009) data, whereas both were included in Xu et al. (2020). In general, the estimated
half-life using the linear mixed effect model were similar to originally reported values in 01 sen
et al. (2009). For instance, as compared with the reported average of 25.8 days ranging from
13.1-45.7 days (01 sen et al.. 2009). the estimated population elimination half-life for serum
PFBS was 25.0 days with individual estimates of 14.6-42.9 days using the linear mixed effect
model.
Although the estimated serum half-lives of PFBS in 01 sen et al. (2009) overlapped with
those of Xu et al. (2020) (mean = 43.8 days, range = 21.9-87.6 days), there is a statistically
significant difference between these two studies as suggested by both parametric (one-way
analysis of variance [ANOVA]) and nonparametric analyses (Kruskal-Wallis test). Overall, the
estimated serum half-life of PFBS by Xu et al. (2020) is about twofold higher than 01 sen et al.
(2009).
Some of the difference between Xu et al. (2020) and 01 sen et al. (2009) may be due to
the difference in initial concentration, where the 01 sen et al. (2009) subjects had initial
concentrations ranging from 100-1,000 ng/mL PFBS, while the highest initial concentrations in
Xu et al. (2020) was 1.3 ng/mL. It is possible that the higher serum levels in the 01 sen et al.
(2009) subjects resulted in saturation of renal resorption, hence more rapid excretion/shorter
half-lives. However, to the extent that some ongoing low-level exposure occurred during the
period of observation, such exposure would cause a greater bias towards over-estimation of the
elimination half-life for the Xu et al. (2020) subjects than those of 01 sen et al. (2009). The data
of 01 sen et al. (2009) might also have a greater signal:noise ratio than the data of Xu et al.
(2020). Despite this uncertainty, the fact that the blood concentrations of the Xu et al. (2020) are
more representative of environmental exposure, that their sample size was larger, and a
significant statistical difference was observed, the two data sets will not be combined and the
half-life estimated by Xu et al. (2020) is presumed to better predict human dosimetry at
environmental levels.
The possibility that menstrual blood loss could contribute to overall clearance was
evaluated, assuming that the concentration of PFBS in menstrual blood is the same as in the
general circulation and that the Vd in humans is equal to the average value estimated for monkeys
(0.23 L/kg). The results indicate that this avenue of loss is more than two orders of magnitude
slower than that indicated by the measured PFBS half-life in humans. Thus, menstrual blood
loss is unlikely to contribute significantly to overall PFBS elimination.
1.3.6 Physiologically Based Pharmacokinetic Models
Fabrega et al. (2015) developed a physiologically based pharmacokinetic model to
estimate the concentration of PFAS, including PFBS, in human tissues based on an existing
model and experimental data on concentrations of PFAS in human tissues from individuals in
Catalonia, Spain. Several uncertainties in the model limit the use for this assessment of PFBS.
There are three chemical-specific parameters that determine the rate of elimination: the
free fraction in blood, the maximum rate of resorption in the kidney (Tm), and the saturation
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constant for that resorption (Xt). No details beyond a rough description are provided on how
these parameter values were identified. The data used for calibration are population samples in
adults, who would essentially be at steady state, and only a single average level of exposure and
corresponding blood concentration are reported, precluding the possibility of evaluating
exposure or concentration dependence. In this situation it is not possible to uniquely identify the
three parameters. This lack of identifiability is likely to be an underlying cause of the extreme
variability in the individual parameter values (among the 11 PFAS evaluated) reported by
Fabrega et al. (2015).
In addition, the rate constant for elimination from the glomerular filtrate compartment to
the urine "storage" compartment (i.e., the bladder) is the total glomerular filtration rate (GFR),
which is approximately 10 L/hour in a 70 kg adult. But most of the glomerular flow is resorbed
in the nephrons, and human urinary output is less than 2 L/day. Hence, the use of GFR for
elimination is not realistic. Finally, note that while the model structure and the equations listed
by Fabrega et al. (2015) appear to be appropriate for most humans, excretion via lactation is not
included.
Of considerable concern is the way in which partition coefficients (PCs) were identified.
In particular, PCs were obtained by taking tissue concentration data from cadavers and
comparing those to average blood concentrations from volunteer subjects, albeit from the same
geographical area (county in Spain). The liver:blood PC for PFDA was thereby estimated to be
0.001 while the value for PFNA was 1.65. By contrast, Kim et al. (2019) obtained values of
-0.6-0.7 for PFDA in male and female rats, -1.2 for PFNA in male rats, and -0.5 for PFNA in
female rats. Thus, there seems to be extreme inconsistency and hence uncertainty in these
parameters as estimated by Fabrega et al. (2015). Generally, human PCs should have values
similar to those in rats.
The study authors do not compare model predictions for Tarragona County, Spain, with
measured values for county residents (i.e., the data used for model calibration). Also, the study
authors state that 20-30 years of simulated time are required to reach steady state. These
steady-state estimates are inconsistent with the elimination data from 01 sen et al. (2009). in
which the half-life in males was 24 days, and in one female subject 46 days. These empirical
half-lives are consistent with a time to steady state of less than a year, indicating that the
predicted clearance from Fabrega et al. (2015) may be an order of magnitude or more too low.
At the same time, the simulated levels of five PFAS (average levels) were consistently lower
than the averages in the validation data, four of these being lower by an order of magnitude or
more.
Thus, predictions of the Fabrega et al. (2015) model are considered highly uncertain, and
data other than those used by the study authors will be needed to accurately estimate key
pharmacokinetic (PK) parameters for PFBS and these other PFAS, a task that would require
significant additional research.
1.3.7 Summary
Collectively, elimination half-lives appear to be similar for mice and rats, with potential
sex-specific toxicokinetic differences being reported (i.e., females appearing to have a faster
elimination rate). Humans have a longer serum elimination half-life (-weeks) than both rodents
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(-hours) and monkeys (-days). Further, although Vd information is not available for humans,
observations in male and female mice, rats, and monkeys exposed to comparable doses indicate
comparability across species. Results of all studies discussed in this section are summarized in
Table 2.
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2.0 PROBLEM FORMULATION
2.1 CONCEPTUAL MODEL
A conceptual model was developed to summarize the availability of data to understand
potential health hazards related to exposure to PFBS and/or K+PFBS. The potential sources of
these chemicals, the routes of exposure for biological receptors of concern (e.g., various human
activities related to ingested drinking water, and food preparation and consumption), the
potential organs and systems affected by exposure (e.g., effects such as developmental toxicity),
and potential populations at risk due to exposure to PFBS and/or potassium salt are depicted in
the conceptual diagram in Figure 2. Arrows indicate linkage between one or more boxes
between levels of organization.
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EPA 690 R-2 J 00 IF
STRESSOR
POTENTIAL
SOURCES OF
EXPOSURE
PFBS and Potassium Salt
Drinking
Water
Ambient
Water
Industrial
Uses
Consumer
Products
Food
Dust
Ail-
Soil
Fire
fighting
Foams
EXPOSURE ROUTES
Oral
Dermal
Inhalation
POTENTL4L
ORGANS/
SYSTEMS
AFFECTED
POTENTIAL
RECEPTORS IN
GENERAL
POPULATION
Thyroid Effects
Reproductive
Effects
Developmental
Effects
Renal Effects
Hepatic Effects
Adults
Children
Pregnant Women
and Fetuses

Lipid and
Lipoprotein
Effects
Lactatlns Women
LEGEND
Data Selected for
Assessment
Limited Data
Unknown
Quantitative Data
Figure 2. Conceptual Model for PFBS and/or Potassium Salt
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2.2 OBJECTIVE
The overall objective of this assessment is to provide the health effects basis for the
development of oral reference doses (RfDs) for PFBS (CASRN 375-73-5) and its related
compound K+PFBS (CASRN 29420-49-3), including the science-based decisions providing the
basis for identifying potential human health effects and estimating PODs. Based on the needs of
the U.S. EPA partner program offices, regions, states, and/or tribes as they pertain to diverse
exposure scenarios and human populations, subchronic and chronic RfDs have been derived.
The assessment includes studies and information previously provided in the 2014 PPRTV
assessment (U.S. EPA. 20140 and builds upon data from the literature published since that
review.
2.3 METHODS
2.3.1 Literature Search
Four online scientific databases (PubMed, Web of Science, TOXLINE, and TSCATS via
TOXLINE) were searched by the U.S. EPA's Health and Environmental Research Online
(HERO) staff and stored in the HERO database.5 The literature search 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. Full details of the search strategy for each
database are presented in Appendix A. The initial database searches were conducted on July 18,
2017 and updated on February 28, 2018; May 1, 2019; and May 15, 2020. Additional studies
[e.g., Lau et al. (2020): Xu et al. (2020)1 were identified during subsequent review periods and
integrated into the assessment as appropriate. Studies were also identified from other sources
relevant to PFBS, including studies submitted to the U.S. EPA by the manufacturer of PFBS
(i.e., 3M) as part of the Toxic Substances Control Act (TSCA) premanufacture notices for other
PFAS chemicals or as required under TSCA reporting requirements and studies referenced in
prior evaluations of PFBS toxicity (MDH. 2020; ATS DR. 2015). In addition, on March 29,
2018, the National Toxicology Program (NTP) published study tables and individual animal data
from a 28-day toxicity study of PFBS
(http://doi.org/10.22427/NTP-DATA-002-01134-0003-0000-4). with a protocol outlining the
NTP study methods available in HERO
(https://hero.epa.gov/hero/index.cfm/reference/details/reference id/4309741) (NTP. 2011). The
final NTP Technical Report on the Toxicity Studies of Perfluoroalkyl Sulfonates Administered by
( lavage to Sprague-Dawley Rats was published in August, 2019 (NTP. 2019).
2.3.2 Screening Process
Two screeners independently conducted a title and abstract screening of the search results
using DistillerSR6 to identify study records that met the Population, Exposure, Comparator, and
Outcome (PECO) eligibility criteria (see Appendix B for a more detailed summary):
5The U.S. EPA's HERO database provides access to the scientific literature behind U.S. EPA science assessments.
The database includes more than 2,500,000 scientific references and data from the peer-reviewed literature used by
the U.S. EPA to develop its regulations.
' DistillerSR is a web-based systematic review software used to screen studies available at
https://www.evidencepartners.com/products/distillersr-svstematic-review-software.
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•	Population: Human and nonhuman mammalian animal species (whole organism) of any
life stage and in vitro models of genotoxicity.
•	Exposure: Any qualitative or quantitative estimates of exposure of PFBS or K+PFBS, via
oral or inhalation routes of exposure. (Note: Nonoral and noninhalation studies are
tracked as potential supplemental material and are presented in Section 4.8.2.)
•	Comparator: A comparison or reference population exposed to lower levels or for shorter
periods of time for humans. Exposure to vehicle-only or untreated control in animals.
•	Outcome: Any examination of cancer or noncancer health outcomes.
In addition to the PECO criteria, the following additional exclusion criteria were applied,
although these study types were tracked as supplemental material as described following the
exclusion criteria:
•	Records that do not contain original data such as other agency assessments, scientific
literature reviews, editorials, and commentaries;
•	Abstract only (e.g., conference abstracts); and
•	Retracted studies.
Records that were not excluded based on title and abstract screening advanced to full-text
review using the same PECO eligibility criteria. Studies that have not undergone peer review
were included if the information could be made public and sufficient details of study methods
and findings were included in the reports. Full-text copies of potentially relevant records
identified from title and abstract screening were retrieved, stored in the HERO database, and
independently assessed by the screeners using DistillerSR to confirm eligibility. At both
title/abstract and full-text review levels, screening conflicts were resolved by discussion between
the primary screeners in consultation with a third reviewer to resolve any remaining
disagreements. During title/abstract or full-text level screening, studies that were not directly
relevant to the PECO, but could provide supplemental information, were categorized (or
"tagged") by the type of supplemental information they provided (e.g., review, commentary, or
letter with no original data; conference abstract; toxicokinetics; mechanistic information aside
from in vitro genotoxicity studies; other routes of exposure; exposure only). Conflict resolution
was not required during the screening process to identify supplemental information (i.e., tagging
by a single screener was sufficient to identify the study as potential supplemental information).
2.3.3 Study Evaluation
Study evaluation was conducted by one reviewer for epidemiological studies and by two
independent reviewers for animal studies using the U.S. EPA's version of Health Assessment
Workspace Collaborative (HAWC), a free and open source web-based software application
designed to manage and facilitate the process of conducting literature assessments.7 For
pragmatic purposes, only one reviewer was considered necessary for epidemiological studies
because it was apparent during literature screening that the animal evidence would be the most
informative for deriving toxicity values. The available outcomes in the epidemiological studies
were heterogeneous and unrelated to each other, and only a single study was available for each
7HAWC: A modular web-based interface to facilitate development of human health assessments of chemicals
(https://hawcBroject.org/).
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outcome. This approach is consistent with recommendations from the National Academies of
Science encouraging the U.S. EPA to explore ways to make systematic review more feasible,
including a "rapid review in which components of the systematic review process are simplified
or omitted (e.g., the need for two independent reviewers)" (NASHM. 2017). Study evaluation
was not conducted for studies tagged during screening as supplemental information.
The general approach for evaluating epidemiology and animal toxicology was the same
(see Figure 3), but the specifics of applying the approach differed. These evaluations were
focused on the methodological approaches and completeness of reporting in the individual
studies, rather than on the direction or magnitude of the study results. Evaluation of
epidemiology studies was conducted for the following domains: exposure measures, outcome
measures, participant selection, confounding, analysis, sensitivity, and selective reporting. For
animal studies, the evaluation process focused on assessing aspects of the study design and
conduct through three broad types of evaluations: reporting quality, risk of bias, and study
sensitivity. A set of domains with accompanying core questions fall under each evaluation type
and directed individual reviewers to evaluate specific study characteristics. For each domain
evaluated for experimental animal studies (reporting quality, selection or performance bias,
confounding/variable control, reporting or attrition bias, exposure methods sensitivity, and
outcome measures and results display), basic considerations provided additional guidance on
how a reviewer might evaluate and judge a study for that domain. Core and prompting questions
used to guide the criteria and judgment for each domain are presented in Appendix C. Key
concerns for the review of epidemiology and animal toxicology studies are potential sources of
bias (factors that could systematically affect the magnitude or direction of an effect in either
direction) and insensitivity (factors that limit the ability of a study to detect a true effect).
For each study in each evaluation domain, reviewers reached a consensus rating
regarding the utility of the study for hazard identification, with categories of good, adequate,
deficient, not reported, or critically deficient. These ratings were then combined across domains
to reach an overall classification of high, medium, or low confidence or uninformative
(definitions of these classifications are available in Appendix C). The rationale for the
classification, including a brief description of any identified strengths and/or limitations from the
domains and their potential impact on the overall confidence determination, is documented and
retrievable in HAWC. Uninformative studies were not used in evidence synthesis or
dose-response analysis. Studies were evaluated for their suitability for each health outcome
investigated and could receive different ratings for each outcome.
For epidemiological studies, exposure-specific criteria were developed prior to evaluation
and are described in detail in Appendix C. In brief, standard analytical methods of measurement
of PFBS in serum or whole-blood using quantitative techniques such as liquid
chromatograph-triple quadrupole mass spectrometry and high-pressure liquid chromatography
with tandem mass spectrometry were preferred. In addition, exposure must have been assessed
in a relevant time window for development of the outcome.
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Individual evaluation domains
\7
Domain judgments
Judgement
Good
Adequate
Deficient
Critically
Deficient
Animal
Epidemiological
Reporting Quality
Exposure measurement
Selection or Performance Bias
Outcome ascertainment
Confounding/Variable Control
Population Selection
Reporting or Attrition Bias
Confounding
Exposure Methods Sensitivity
Analysis
Outcome Measures and Results
Display
Sensitivity
Selective reporting
Interpretation
Appropriate study conduct relating to the domain & minor
deficiencies not expected to influence results.
A study that may have some limitations relating to the domain, but
they are not likely to be severe or to have a notable impact on
results.
Identified biases or deficiencies interpreted as likely to have had a
notable impact on the results or prevent reliable interpretation of
study findings.
A serious flaw identified that is interpreted to be the primary driver
of any observed effect or makes the study uninterpretable. Study is
not used without exceptional justification.
Overall study rating
for an outcome
Rating
Interpretation
High
No notable deficiencies or concerns identified; potential for bias unlikely
or minimal: sensitive methodology.
Medium
Possible deficiencies or concerns noted, but resulting bias or lack of
sensitivity would be unlikely to be of a notable degree.
Low
Uninformative
Deficiencies or concerns were noted, and the potential for substantive
bias or inadequate sensitivity could have a significant impact on the study
results or their interpretation.
Serious flaw(s) makes study results unusable for hazard identification.
Figure 3. Approach for Evaluating Epidemiological and Animal Toxicology Studies
2.3.4 Data Extraction
Information on study design, methods, results, and data from animal toxicology studies
were extracted into the HAWC and are available at
https://hawcprd.epa.gov/assessment/100000037/. Visual graphics prepared from HAWC are
embedded as hyperlinks and are fully interactive when viewed online by way of a "click to see
more" capability. Clicking on content allows access to study evaluation ratings, methodological
details, and underlying study data. The action of clicking on content contained in those visual
graphics (e.g., data points, endpoint, and study design) will yield the underlying data supporting
the visual content.8 A HAWC user guide can be found in Appendix D. Study methods and
findings from epidemiological studies were described in narratives, given the small size and
heterogeneity of the evidence base. Data extraction was performed by one member of the
evaluation team and checked by one to two other members. Any discrepancies in data extraction
8The following browsers are fully supported for accessing HAWC: Google Chrome (preferred), Mozilla Firefox, and
Apple Safari. There are errors in functionality when viewed with Internet Explorer.
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were resolved by discussion or consultation with a third member of the evaluation team. Digital
rulers such as WebPlotDigitizer and Grab It! (https://automeris.io/WebPlotDigitizer/ and
https://grab-it.softl 12.com/. respectively) were used to extract numerical information from
figures. Use of digital rulers was documented during extraction. Dose levels were extracted as
reported in the study and converted to mg/kg-day (HED) for endpoints that were considered for
use in the dose-response and derivation of toxicity values.
2.3.5 Evidence Synthesis
For the purposes of this assessment, after study evaluation, the informative evidence for
each outcome was summarized from the available human studies and, separately, the available
animal studies. This synthesis provides a short synopsis of the breadth of data available to
inform each outcome and summarizes information on the general study design, doses tested,
outcomes evaluated, and results for the endpoints of interest within each study. While the
evidence synthesis describes inferences about the methodological rigor and sensitivity of the
individual studies (i.e., study confidence) and discusses the pattern and magnitude of the
experimental findings within studies, it does not include conclusions drawn across the sets of
studies (see "Evidence Integration and Hazard Characterization," next).
2.3.6 Evidence Integration and Hazard Characterization
In this assessment, the evaluation of the available evidence from informative human and
animal studies was described in an evidence integration narrative for each outcome, including
overall evidence integration judgments as to whether the data provide evidence sufficient to
support a hazard. These integrated judgments serve to characterize the extent of the available
evidence for each outcome, including information on potential susceptible populations and life
stages, as well as important uncertainties in interpreting the data.
The evidence integration for each health effect considered aspects of an association that
might suggest causation first introduced by Austin Bradford Hill (Hill. 1965). including the
consistency, exposure-response relationship, strength of association, biological plausibility, and
coherence of the evidence. This involved weighing the PFBS-specific human and animal
evidence relating to each of these considerations within or across studies, including both
evidence that supports causation as well as evidence that indicates lack of support. For example,
the evaluation of consistency examined the similarity of results across studies (e.g., direction and
magnitude). When inconsistencies across studies were identified, the evaluation considered
whether results were "conflicting" (i.e., unexplained positive and negative results in similarly
exposed human populations or in similar animal models) or "differing" (i.e., mixed results
explained by differences between human populations, animal models, exposure conditions, or
study methods), based on analyses of potentially important explanatory factors such as
confidence in the studies' results (the results of higher confidence studies were emphasized),
exposure levels or duration, or differences in populations or species (including potential
susceptible groups) across studies (U.S. EPA. 2005). While consistent evidence across studies
increases support for a hazard, unexplained inconsistency or conflicting evidence decreases
support for a hazard. The evaluations of these considerations were informed by U.S. EPA
guidelines, including Guidelines for Developmental Toxicity Risk Assessment (U.S. EPA, 1991a)
and Guidelines for Reproductive Toxicity Risk Assessment (U.S. EPA. 1996b).
25

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EPA/690/R-21/001F
The overall evidence integration judgments were developed using a structured framework
based on evaluation of the considerations above (see Table 3). Using this framework, the human
and animal evidence for each health effect was judged separately as supports a hazard,
equivocal, or supports no hazard. Evidence integration judgments of supports a hazard span a
range of supportive evidence bases that can be further differentiated by the quantity and quality
of information available to rule out alternative explanations for the results. Equivocal evidence
is limited in terms of the quantity, consistency, or confidence level of the available studies and
serves to encourage additional research. Supports no hazard requires several high-confidence
studies across potentially susceptible populations with consistent null results; this judgment was
not reached in this assessment. Overall evidence integration judgments were made based on
conclusions from both the animal and human data, considering the available information on the
human relevance of findings in animals. Thus, for example, evidence in animals that supports a
hazard alongside equivocal human evidence in the absence of information indicating that the
responses in animals are unlikely to be relevant to humans would result in an overall judgment of
supports a hazard for that outcome.
Table 3. Criteria for Overall Evidence Integration Judgments

Animal
Human
Supports a
hazard
The evidence for effects is consistent or largely
consistent in at least one high- or
medium-confidence experiment.3 Although
notable uncertainties across studies might remain,
any inconsistent evidence or remaining
uncertainties are insufficient to discount the cause
for concern from the positive experiments. In the
strongest scenarios, the set of experiments provide
evidence supporting a causal association across
independent laboratories or species. In other
scenarios, including evidence for an effect in a
single study, the experiment(s) demonstrate
additional support for causality such as coherent
effects across multiple related endpoints; an
unusual magnitude of effect, rarity, age at onset,
or severity; a strong dose-response relationship;
and/or consistent observations across exposure
scenarios (e.g., route, timing, or duration), sexes,
or animal strains.
One or more high- or medium-confidence
independent studies reporting an association
between the exposure and the health outcome. In
general, the study results are largely consistent or
any inconsistent results are insufficient to
discount the cause for concern from the higher
confidence study or studies, and there is
reasonable confidence that alternative
explanations, including chance, bias, and
confounding, have been ruled out. In situations in
which only a single study is available, the results
of multiple studies are heterogeneous, or
alternative explanations, including chance, bias
and confounding, have not been ruled out, there is
additional supporting evidence such as
associations with biologically related endpoints in
other human studies (coherence), large estimates
of risk, or strong evidence of an
exposure-response within or across studies.
26

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EPA/690/R-21/001F
Table 3. Criteria for Overall Evidence Integration Judgments

Animal
Human
Equivocal
The evidence is generally inadequate to determine
hazard. This includes a lack of relevant studies
available or a set of low-confidence experiments.
It also includes scenarios with a set of high- or
medium-confidence experiments that are not
reasonably consistent or not considered
informative to the hazard question under
evaluation. This category would also include a
single high- or medium-confidence experiment
with weak evidence of an effect (e.g., changes in
one endpoint among several related endpoints,
and without additional evidence supporting
causality).
The evidence is considered inadequate to describe
an association between exposure and the health
outcome with confidence. This includes a lack of
studies available in humans, only low-confidence
studies, or considerable heterogeneity across
medium- or high-confidence studies. This also
includes scenarios in which there are serious
residual uncertainties across studies (these
uncertainties typically relate to exposure
characterization or outcome ascertainment,
including temporality) in a set of largely
consistent medium- or high-confidence studies.
Supports no
hazard
A set of high-confidence experiments examining
the full spectrum of related endpoints within a
type of toxicity, with multiple species, and testing
a reasonable range of exposure levels and
adequate sample size in both sexes, with none
showing any indication of effects. The data are
compelling in that the experiments have examined
the range of scenarios across which health effects
in animals could be observed, and an alternative
explanation (e.g., inadequately controlled features
of the studies' experimental designs) for the
observed lack of effects is not available. The
experiments were designed to specifically test for
effects of interest, including suitable exposure
timing and duration, post-exposure latency, and
endpoint evaluation procedures, and to address
potentially susceptible populations and life stages.
Several high-confidence studies, showing
consistently null results (e.g., an OR of 1.0) ruling
out alternative explanations including chance,
bias, and confounding with reasonable
confidence. Each of the studies should have used
an optimal outcome and exposure assessment and
adequate sample size (specifically for higher
exposure groups and for sensitive populations).
The set as a whole should include the full range of
levels of exposures that human beings are known
to encounter, an evaluation of an exposure
response gradient, and at-risk populations and life
stages and should be mutually consistent in not
showing any indication of effect at any level of
exposure.
""'Experiment" refers to measurements in a single population of exposed animals (e.g., a study that included
separate evaluations of rats and of mice, or separate cohorts exposed at different life stages, would be considered as
multiple experiments). Conversely, two papers or studies that report on the same cohort of exposed animals
(e.g., examining different endpoints) would not be considered separate experiments.
OR = odds ratio.
The primary evidence and rationale supporting these decisions were summarized in a
single evidence profile table to transparently convey the aspects of the evidence that were
considered to increase or decrease the hazard support for each health effect. For the purposes of
this assessment, only the integrated evidence that supports a hazard was considered for use in
the dose-response analysis and derivation of toxicity values.
2.3.7 Derivation of Values
Development of the dose-response assessment for PFBS and/or the potassium salt has
followed the general guidelines for risk assessment put forth by the National Research Council
(NRC. 1983) and the U.S. EPA's h'ramework for Human Health Risk Assessment to Inform
27

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EPA/690/R-21/001F
Decision Making (U.S. EPA. 2014c). Other U.S. EPA guidelines and reviews considered in the
development of this assessment include the following:
• A Review of the Reference Dose and Reference Concentration Processes (U.S. EPA.
2002).
• A Framework for Assessing Health Risks of Environmental Exposures to Children (U.S.
HP A. 2006).
• Exposure Factors Handbook (U.S. EPA. 201 la).9
• Recommended Use of Body Weight3/4 as the Default Method in Derivation of the Oral
Reference Dose (U.S. EPA. 201 lb).
• Guidance for Applying Quantitative Data to Develop Data-Derived Extrapolation
Factors for Interspecies and Intraspecies Extrapolation (U.S. EPA. 2014d).
• Benchmark Dose Technical Guidance Document (U.S. EPA. 2012).
• Child-Specific Exposure Scenarios Examples (U.S. EPA. 2014a).
The U.S. EPA's A Review of the Reference Dose and Reference Concentration Processes
describes a multistep approach to dose-response assessment, including analysis in the range of
observation followed by extrapolation to lower levels (U.S. EPA. 2002). As described above,
before deriving toxicity values, the U.S. EPA conducted a comprehensive evaluation of available
human epidemiological and animal toxicity studies to identify potential health hazards and
associated dose-response information through the literature search and screening, study
evaluation, evidence synthesis, and evidence integration steps. This evaluation informed the
selection of candidate key studies and critical effects for dose-response analysis, from which the
U.S. EPA identified a critical effect and point of departure (POD) for subchronic and chronic
reference value derivation and extrapolated a selected POD to a corresponding RfD
(e.g., subchronic RfD). For dose-response analysis of PFBS and/or the potassium salt, the
U.S. EPA used the BMD approach to identify a POD. The steps for deriving an RfD using the
BMD approach are summarized below.
• Step 1: Evaluate the data to identify and characterize endpoints related to exposure
to PFBS chemicals. This step involved determining the relevant studies and adverse
effects to be considered for BMD modeling. Once the appropriate data were collected,
evaluated for study quality, and characterized for adverse outcomes, endpoints were
selected that were judged to be relevant (i.e., for the purposes of this assessment, effects
that were sufficient to support a hazard) and sensitive as a function of dose (typically
defined by the no-observed-adverse-effect level [NOAEL] value). In this assessment,
these decisions were directly informed by the evidence integration judgments arrived at
for each assessed health outcome. Some of the most important considerations that
influenced selection of endpoints for BMD modeling include data showing a dose-
response relationship, percent change from controls, adversity of effect, and consistency
across studies. For PFBS, thyroid, developmental, and kidney endpoints were considered
for toxicity value derivations.
9Please note that specific updates to this handbook are available at
https://cfpub.epa. gov/ncea/risk/recordisplay.cfm?deid=236252.
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EPA/690/R-21/001F
• Step 2: Convert the adjusted daily doses to an HED. The adjusted daily doses were
converted to HEDs by considering U.S. EPA's Recommended Use of Body Weight3/4 as
the Default Method in Derivation of the Oral Reference Dose (U.S. EPA. 201 lb).
• Step 3: Select the benchmark response (BMR) level. The endpoints selected were
modeled using the U.S. EPA's Benchmark Dose Technical Guidance Document (U.S.
EPA. 2012). The BMR is a predetermined change in the response rate of an adverse
effect. It serves as the basis for obtaining the benchmark dose lower confidence limit
(BMDL), which is the 95% lower bound of the BMD. BMRs were identified and applied
consistent with quantal and continuous data and, when possible, informed by
understanding of biological significance.
• Step 4: BMD model the data. This step involved fitting a statistical model to the
dose-response data that describes the data set of the identified adverse effect. Typically,
this involved selecting a family or families of models (e.g., polynomial continuous, Hill
continuous, or exponential continuous) for further consideration based on the data and
experimental design. In this step, a BMDL was derived by placing confidence limits
(one- or two-sided) and a confidence level (typically 95%) on a BMD to obtain the dose
that ensures with high confidence that the BMR is not exceeded.
• Step 5: Determine a POD (HED). If modeling was feasible, the estimated BMDL
(HED)s were used as PODs (i.e., POD [HED]). If dose-response modeling was not
feasible, NOAEL (HED)s or lowest-observed-adverse-effect level (LOAEL) (HED)s
were identified.
• Step 6: Provide rationale for selecting uncertainty factors. Uncertainty factors were
selected in accordance with U.S. EPA guidelines considering variations in sensitivity
among humans, differences between animals and humans, the duration of exposure in the
key study compared to a lifetime of the species studied, and the potential limitations of
the toxicology database (U.S. EPA. 2014d. 2011b. 2002. 1994).
• Step 7: Calculate the subchronic and chronic RfDs. The RfDs were calculated by
dividing a POD (HED) by the selected uncertainty factors.
RfD = POD (HED)
UFc
where:
POD (HED) is calculated from the BMDL or NOAEL using a BW3 4 allometric scaling
approach consistent with U.S. EPA guidance (U.S. EPA. 201 lb)
UFc is established in accordance with U.S. EPA guidelines (U.S. EPA. 2014d. 2011b.
2002. 1994) considering variations in sensitivity among humans, differences between
animals and humans, the duration of exposure in the key study compared to a lifetime of
the species studied, and the potential limitations of the toxicology database.
• Step 8: Assignment of Confidence Levels. In assessments in which an RfD or RfC is
derived, characterization of the level of confidence in the principal study(ies), the
database associated with that reference value, and the overall confidence in the reference
value(s) are provided. Details on characterizing confidence are provided in Chapter 4
29

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EPA/690/R-21/001F
(specifically Section 4.3.9.2) of the U.S. EPA's Methods for Derivation of Inhalation
Reference Concentrations and Application of Inhalation Dosimetry (U.S. EPA. 1994).
For example, the ranking of confidence in the database (low, medium, or high) reflects
EPA's assessment of the degree to which the reference value (e.g., RfD) might
potentially change (in either direction) with the acquisition of new data.
3.0 OVERVIEW OF EVIDENCE IDENTIFICATION FOR SYNTHESIS AND
DOSE-RESPONSE ANALYSIS
3.1 LITERATURE SEARCH AND SCREENING RESULTS
The database searches yielded 451 unique records, with 50 records identified from
additional sources, such as TSCA submissions, posted NTP study tables, peer-review
recommendations, and review of reference lists from other authoritative sources. Of the
501 studies identified, 377 were excluded during title and abstract screening, 124 were reviewed
at the full-text level, and 42 were considered relevant to the PECO eligibility criteria (see Figure
4). This included 19 epidemiologic studies (described in 22 publications), 10 in vivo animal
studies (described in 15 peer-reviewed and non-peer-reviewed publications), and 5 in vitro
genotoxicity studies. The detailed search approach, including the query strings and PECO
criteria, is provided in Appendix A and Appendix B, respectively.
30

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EPA 690 R-21 00IF
Figure 4. Literature Search and Screening Flow Diagram for PFBS (CASRN 375-73-5)
3.2 STUDY EVALUATION RESULTS
Based on the study evaluations, seven human epidemiology studies were considered
uninformative and are not discussed any further in this assessment (see Table 4). All animal
studies were considered informative and thus were identified as relevant during literature
screening and included in the evidence synthesis and dose-response analysis. Overall,
12 epidemiologic studies (described in 15 publications) and 10 in vivo animal studies (described
in 15 peer-reviewed and non-peer-reviewed publications) were included in the evidence
synthesis and further evaluated for use in the development of toxicity values for PFBS. As
shown in Figures 5 and 6, while the database of studies on PFBS is not large, several high- and
medium-confidence oral exposure studies in animals were identified, as were several
31

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EPA/690/R-21/001F
medium-confidence studies in humans. Multiple publications of the same study are not listed as
independent studies in HAWC, they are reviewed together in one entry. In addition, Shiuc
(2016) was not evaluated because the outcome (i.e., sleep disturbances) was considered a
nonspecific effect, and thus was not entered into HAWC. No studies were identified evaluating
the toxicity of PFBS or K+PFBS following inhalation exposure or on the carcinogenicity of
PFBS or K+PFBS in humans or animals.
Table 4. Epidemiological Studies Excluded Based on Study Evaluation
Reference
Outcome
Reason for Exclusion
Bao et al. (2017)
Blood pressure
Extremely poor sensitivity (96% of participants below the
LOD for PFBS measurement) with no observed
association.
Berk et al. (2014)
Depression
Serious concerns with temporality between exposure and
outcome, confounding, and analysis.
Gvllenhammar et al. (2018)
Birth size, weight gain
Extremely poor sensitivity (median exposure = 0.01 ng/g,
IQR LOD-0.04, 43% below the LOD for PFBS
measurement) with no observed association.

Kim et al. (2016)
Congenital
hypothyroidism
Excluded from full statistical analysis by study authors
because of a high percentage below the LOD (72%) for
PFBS measurement.
Seo et al. (2018)
Cholesterol, uric acid,
diabetes, BMI, thyroid
hormones
No consideration of potential confounding.
Shiue (2016)a
Sleep disturbances
Not evaluated because of nonspecific effect.
Wane et al. (2017)
Endometriosis-related
infertility
Exposure measured concurrent with outcome for chronic
outcome; serious concerns for exposure and outcome
misclassification.

aShiue (2016) was not evaluated because the outcome was sleep disturbances, which was considered a nonspecific
effect, and thus was not entered in HAWC.
BMI = body mass index; HAWC = Health Assessment Workspace Collaborative; IQR = interquartile range;
LOD = limit of detection; PFBS = perfluorobutane sulfonic acid.
32

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EPA 690 R-21 00 IP
&
&
,o°
„0\1 'oO^'oO^A e^®aCV\^^Ve^eooV&'nV°'^-'a'V^'a0^°'l<^ '
Participant selection
Exposure measurement
Outcome ascertainment
Confounding -
Analysis
Sensitivity
Selective Reporting
Overall confidence







Not assessed

Critically deficient
(metric) or
uninfonnative
(overall)

Good (metric)

Adequate (metric)

Deficient

because of

++
or high
+
or medium
-
(metric) or low
N/A
critical


(overall)

(overall)

(overall)

deficiency in
other domain

Figure 5. Evaluation Results for Epidemiological Studies Assessing Effects of PFBS
(Click to see interactive data graphic for rating rationales)
33

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EPA 690 R-21 00 IP

Reporting -
+ +
+ +
+ +
++
++
++
++
++
++
++
Allocation -
+ +
+ +
D

++
++
++
++
++
++
Blinding -



++
++
++

+
Variable Control -
+ +
+ +
++
++
++
++
++
++
++
++
Selective Reporting & Attrition -
+ +
+ +
B

++
++
D
++
++
++
Exposure Characterization -
+ +
+ +
++
D
++
++
++
++
++
++
Utility of Study Design -
+ +
+ +
++
++
++
++
++
++
++
++
Outcome Assessment -
+
+ +
~
++
++
++
++
++
++
++
Results Presentation -
+ +
+ +
++
++
++
++
++
++
++
++
Overall confidence

Good (metric)

Adequate (metric)

Deficient

Not

Critically deficient
++
or high
+
or medium
-
(metric) or low
NR
reported

(metric) or uninfonnative

(overall)

(overall)

(overall)

for metric

(overall)
Figure 6. Evaluation Results for Animal Studies Assessing Effects of PFBS Exposure
(Click to see interactive data graphic for rating rationales)
34

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EPA/690/R-21/001F
4.0 EVIDENCE SYNTHESIS: OVERVIEW OF INCLUDED STUDIES
The database of all repeated-dose oral toxicity studies for PFBS and the related
compound K+PFBS that are potentially relevant for deriving RfD values includes a short-term
range-finding study in rats (3M, 2000d). two 28-day studies in rats (NTP. 2019; 3M, 2001). one
subchronic study in rats (l.ieder et al.. 2009a; York, 2003b). one subchronic-duration lipoprotein
metabolism study in mice (Bijland et al.. 201 1; 3M, 2010). three gestational exposure studies in
mice and rats (Feng et al.. 2017; York. 2003a. 2002). and a two-generation reproductive toxicity
study in rats (l.ieder et al.. 2009b; York. 2003c. d, e). In addition, 19 epidemiological studies
(described in 22 publications) were identified that report on the association between PFBS and
human health effects. Specific study limitations identified during evaluation (see HAWC) are
discussed only for studies interpreted as low confidence or if a limitation affected a specific
inference for drawing conclusions.
Human and animal studies have evaluated potential effects on the thyroid, reproductive
systems, development, kidneys, liver, and lipid and lipoprotein homeostasis following exposure
to PFBS. The evidence base for these outcomes is presented in this section. For each potential
health effect, the synthesis describes the database of human and animal studies, as well as an
array of the animal results across studies. NOAELs and LOAELs presented in the figures and
text are based on statistical significance and/or biological significance (e.g., directionality of
effect [statistically significantly decreased cholesterol/triglycerides is of unclear toxicological
relevance], abnormal or irregular dose-response relationship [nonmonotonicity], tissue-specific
considerations for magnitude of effect [statistically nonsignificant increase of >10% in liver
weight interpreted as biologically significant]). A summary of the available database is
presented in Table 6 of Section 5. For information in this section, evidence to inform
organ/system-specific effects of PFBS in animals following developmental exposure is discussed
in the individual organ/sy stem-specific sections (e.g., reproductive cycling endpoints after
developmental exposure are discussed in the "Reproductive Effects" section). Other effects
informing potential developmental effects (e.g., pup BW) are discussed in the "Offspring Growth
and Early Development" section.
Evidence integration analyses and overall judgments on the hazard support for each
outcome domain provided by the available human and animal studies are discussed in the
"Evidence Integration and Hazard Characterization" section. Notably, in that section, the
evidence informing organ/sy stem-specific endpoints after developmental exposure was
considered potentially informative to both the developmental effects outcome domain and the
organ/sy stem-specific outcome domain.
4.1 THYROID EFFECTS
4.1.1 Human Studies
One low-confidence study examined cross-sectional associations between PFBS exposure
and thyroid hormones in women with premature ovarian insufficiency (Zhang et al.. 2018) and
reported no association with free T3, free T4, or thyroid-stimulating hormone. However, this
study had poor sensitivity and methodological limitations that make interpreting these null
results difficult; further, the results in this highly selected population may not be generalizable.
35

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EPA 690 R-21 00IF
4.1.2 Animal Studies
Two high-confidence studies evaluated the effects of PFBS exposure on the thyroid,
specifically thyroid hormone levels, thyroid histopathology, and thyroid weight (NTP. 2019;
Feng et al.. 2017) (see Figure 7). Dams exposed to K+PFBS through gestation (GDs 1-20)
exhibited a statistically significant decrease in total triiodothyronine (T3). total thyroxine (T4).
and free T4 (reduced 17, 21, and 12%, respectively, relative to control at 200 mg/kg-day and
reduced 16, 20, and 11%, respectively, relative to control at 500 mg/kg-day) on GD 20 at doses
of 200 and 500 mg/kg-day, but not at 50 mg/kg-day (Feng et al.. 2017). Decreased total T3 and
total T4 were also reported at PNDs 1, 30, and 60 in offspring gestationally exposed to K+PFBS
at the same doses (up to 37% reduction in T3 and 52% reduction in T4). Increased
thyroid-stimulating hormone (TSH) was reported in dams and pubertal (PND 30) offspring
(21 and 14% relative to control at 200 mg/kg-day, respectively) exposed gestationally to
K+PFBS. Statistically significant dose-dependent decreases in total T3. total T4. and free T4 were
also reported after exposure in male and female rats to K+PFBS for 28 days at all doses tested
(>62.6 mg/kg-day) (NTP. 2019). The reported reductions in rat total T3 were up to -57% (male)
and -43% (female), in free T4 up to -86% (male) and -77% (female), and in total T4 up to -97%
(male) and -71% (female). Dose-response graphics for T4, T3, and TSH, including effect size
and variability, are included in Appendix E, Figures Figure E-l, Figure E-2, and Figure E-3,
respectively. Thyroid gland weight, thyroid histopathology, and TSH levels were not changed
after 28 days of PFBS exposure in male or female rats at doses up to 1,000 mg/kg-day (NTP.
2019).
ICndpoint Name
.Study Name
Study Type
Animal Description
Observation Time


PFBS Thyroid Effeels
Tetraiodothyroninc (T4), Free
NTP 2018, 4309741
short-term (28 days)
Rat. Harlan Sprague-Dawlev ((f)
Day 28


	v	•
• Doses
Ral. Harlan Spraguc-Dawlcy (9) Day 28

-w-
	~	•
| Dose Range

Feng 2017. 3856465
developmental (GDI to 20)
P0 Mouse. ICR (9)
GD20
M	
^—
	V
A Significant Increase
Tetraiodothyronine (T4), Total
NTP 2018,4309741
short-term (28 days)
Rat, Harlan Sprague-Dawlev (cf)
Day 28

—V-
	~	•
^ Significant Decrease



Rat, Harlan Sprague-Dawlev (9)
Day 28

—V-
	=?	•


Feng 2017, 3856465
developmental (GDI to 20)
PO Mouse, ICR (9)
GD20
~-«	
JS7—
	w




F1 Mouse. ICR (9)
PND1
M	
^—
	~





PND30
	
—
	V





PND60
	
^—
	~


NTP 2018, 4309741
short-term (28 days)
Rat, Harlan Sprague-Dawley (cf)
Day 28












Rat. Harlan Sprague-Dawley (9)
Day 28

~
	~	~


Feng 2017, 3856465
developmental (GDI to 20)
P0 Mouse, ICR (9)
GD20
M	
~—
	V




F1 Mouse, ICR (9)
PND1
	
-v—
	V





PND30
	
^—
	V





PNDfiO
M	
—
	w

Thyroid Stimulating Hormone (TSH)
NTP 2018. 4309741
short-term (28 days)
Rat. Harlan Spraguc-Dawlcy (cf)
Day 28
~ • •	•	•	~




Rat. Harlan Sprague-Dawley (9)
Day 28
~ » »




Feng 2017. 3856465
developmental (GDI to 20)
P0 Mouse. ICR (9)
GD20
M	
-£s—
	A














PND30
	
-£i—
	A





PND60
~-«	



Thyroid Weight. Absolute
NTP 2018, 4309741
short-term (28 days)
Rat. Harlan Spraguc-Dawlcy (cf)
Day 28
~ • •

	•	~




Rat, Harlan Sprague-Dawley (9)
Day 28
» • •

	• ~

Thyroid Weight. Relative
NTP 2018,4309741
short-term (28 days)
Rat. Harlan Sprague-Dawley (cf)
Day 28
» • •

	•	•




Rat, Harlan Sprague-Dawley (9)
Day 28
~ » »

	* ~

Thyroid Histopathology
NTP 2018, 4309741
short-term (28 days)
Rat. Harlan Sprague-Dawley (cf)
Day 28
» • •

	•	~




Rat, Harlan Sprague-Dawley (9)
Day 28
~ • •	•	•	~


3M, 2000, 4289992
short-term (10 days)
Rat, Crl: Cd (Sd) lbs Br (cf)
Day 11
~	•—

	~




Rat, Crl: Cd (Sd) lbs Br (9)
Day 11
~	•—

	~





-1
X) 0 100
200 X
400 500 600 700 800 900 1.000 1.
00







Dose (mg/kg/day)
Figure 7. Thyroid Effects from K+PFBS Exposure
(Click to see interactive data graphic and rationale for study evaluations for effects on the
thyroid in HAWC)
36

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EPA/690/R-21/001F
4.2 REPRODUCTIVE EFFECTS
4.2.1	Human Studies
Five studies of populations in China and Taiwan examined different reproductive
outcomes in women and men (Yao et al.. 2019; Song et al.. 2018; Zhang et al.. 2018; Zhou et al..
2017a; Zhou et al.. 2016).
Three low-confidence studies examined reproductive hormones in newborn boys and
girls in China (Yao et al.. 2019). adolescent boys and girls in Taiwan (Zhou et al.. 2016). and
adult women in China (Zhang et al.. 2018). The study in newborns reported lower testosterone
(P: -0.23; 95% CI: -0.46-0.01) and estradiol (P: -0.09; 95% CI: -0.2-0.01) in cord blood in
male babies, but these differences were not statistically significant (Yao et al.. 2019). The other
two studies reported no clear associations between PFBS levels and reproductive hormones in
women with premature ovarian insufficiency (Zhang et al.. 2018) or in adolescents, either among
the entire study population or stratified by sex (Zhou et al.. 2016).
One low-confidence cross-sectional study (Song et al.. 2018) examined the association
between PFBS exposure and semen parameters. There was no indication of decreased semen
quality in this study (correlation coefficients of-0.022 for semen concentration and 0.195
\p < 0.05] for progressive motility), although issues were noted regarding the ability of this study
to detect an effect and important methodological details were missing.
Two studies examined other female reproductive effects: a cross-sectional study of
menstrual cycle characteristics in a general population sample of women planning to become
pregnant who were enrolled at preconception care clinics in China (Zhou et al.. 2017a) and a
case-control study in China of premature ovarian insufficiency (Zhang et al .. 2018). defined by
FSH level and oligo/amenorrhea. For any outcome related to menstruation, there is significant
potential for reverse causation because menstruation is a potential mechanism by which PFAS
are removed from the body (Wong et al .. 2014; Zhang et al .. 2013); therefore, both of these
studies are considered low confidence. Zhou et al. (2017a) reported adjusted odds ratios (OR) of
1.30 (95% CI: 0.54-3.12) for menorrhagia and 1.48 (95% CI: 0.54-4.03) for hypomenorrhea in
preconception women in China for each one unit increase in PFBS, but these results were not
statistically significant. The study authors also reported inverse statistically nonsignificant
associations for these two outcomes based on exposure quartiles (OR range: 0.61-0.84 for the
highest quartiles relative to the referent) with no evidence of an exposure-response relationship,
indicating that the associations are not robust. All of the analyses in this study examined
continuous outcome measures. Zhang et al. (2018) reported no increase in odds of premature
ovarian insufficiency with higher PFBS exposure (OR for tertile 2 vs. tertile 1: 0.84, 95% CI:
0.44-1.60; OR for tertile 3: 0.92, 95% CI: 0.48-1.76).
4.2.2	Animal Studies
Reproductive outcomes were evaluated in a high-confidence study of prenatal exposure
to PFBS in mice (Feng et al.. 2017). in two high-confidence gestational exposure studies in rats
(York. 2003c. 2002). in high-confidence short-term and sub chronic studies in rats fNTP (2019)
and l.ieder et al. (2009a). respectively], and in a high-confidence two-generation reproductive
study in rats (l.ieder et al.. 2009b). Endpoints evaluated in these studies include fertility and
37

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EPA/690/R-21/001F
pregnancy outcomes, hormone levels, markers of reproductive development, and reproductive
organ weights.
4.2.2.1 Female Fertility and Pregnancy Outcomes
Female fertility parameters were evaluated by both Feng et al. (2017) and Lieder et al.
(2009b). who reported generally no effects in exposed parents, but some effects after gestational
exposure in the Fi offspring (click to see interactive graphic for female fertility effects in
HAWC). Female fertility (e.g., fertility index and days in cohabitation) and delivery parameters
(e.g., length of gestation, % deliveries, stillborn pups, and implantation sites) evaluated in l.ieder
et al. (2009b) were generally unaffected by K PFBS treatment for Po- and Fi-generation dams at
doses up to 1,000 mg/kg-day. The mean number of live born Fi pups was statistically
significantly decreased in the 30-mg/kg-day group, but this change was not dose dependent. The
viability index in Fi pups and the lactation index in Fi and F2 pups showed statistically
significant changes at various doses but were not dose dependent (l.ieder et al.. 2009b).
Similarly, no effects were observed in delivery and litter parameters (e.g., implantations, litter
sizes, live fetuses, corpora lutea, and early resorptions) following prenatal exposure from GDs 6
to 20 (York. 2003c. 2002). Adult (PND 60) Fi females gestationally exposed to PFBS at doses
>200 mg/kg-day, however, exhibited fewer primordial, primary, secondary, early antral, antral,
and preovulatory follicles, as well as fewer corpora lutea than control animals (Feng et al.. 2017).
Importantly, no effects on the health (e.g., weight gain) of the exposed dams were observed at
any dose (Feng et al.. 2017). l.ieder et al. (2009b) evaluated ovarian follicles in Fi females after
they were mated and their pups had been weaned (i.e., Lactation Day [LD] 22) and observed no
effects compared with controls at 1,000 mg/kg-day; however, no quantitative data were reported.
Ovarian parameters were not evaluated in the study by York (2002).
4.2.2.2 Male Fertility
Two studies using S-D rats evaluated several potential responses in the male reproductive
system (N I P. 2019; l.ieder et al.. 2009b). Male fertility parameters and reproductive effects
(e.g., sperm parameters) were generally unaffected by K+PFBS treatment in Po- and
Fi-generation males observed by l.ieder et al. (2009b). At the highest dose, there were
statistically significant increases in the percentage of abnormal sperm in Fi animals and
decreases in testicular sperm count in Po-generation males. In addition, the study authors
reported that the number of spermatids per gram testis was within the historical control of the
testing facility. These effects were not statistically changed at lower doses. Alterations in
parameters such as sperm count/number and morphology are considered indicative of adverse
responses in the male reproductive system (Foster and Gray. 2013; Mangelsdorf et al.. 2003;
U.S. EPA. 1996a). A 28-day exposure study reported a decreased trend in testicular spermatid
count per mg testis evaluated at the time of necropsy; however, no significant effects on other
sperm measures were reported, including caudal epididymal sperm count and sperm motility
(N I P. 2019). Note that a complete spermatogenesis cycle in male rats is typically 7 weeks in
length, thus study designs of shorter duration could potentially miss effects of chemical exposure
on some sperm parameters. Accordingly, the differences in responses observed in the two
available studies might have been due to experimental design differences, because l.ieder et al.
(2009b) exposed Po animals for 70 days and Fi animals during the entire period of gestation plus
lactation, whereas N'l'P (2019) exposed animals for 28 days. Future studies should be conducted
38

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EPA 690 R-21 00IF
to determine whether long-term and/or gestational exposure to PFBS significantly affects sperm
measures in sexually mature and developing animals.
4.2.2.3 Reproductive Hormones (Female and Male)
Reproductive hormones were evaluated in mice (Feng et al.. 2017) and, to a limited
extent, in rats (NTP. 2019) (see Figure 8). Exposure to K+PFBS for 28 days resulted in a
significant trend for increased testosterone levels in females, but not in males (NTP. 2019). The
increase in testosterone was not statistically significant when compared to control at any dose by
pairwise analysis. Prenatal exposure to PFBS at and above 200 mg/kg-day resulted in
statistically significant reduced serum estradiol levels and increased serum luteinizing hormone
levels in pubertal offspring (i.e., PND 30) (Feng et al.. 2017). The change in serum estradiol
levels, but not luteinizing hormone, continued into adulthood in the K+PFBS-exposed offspring
(i.e., PND 60). Adult PFBS-exposed offspring also exhibited decreased serum progesterone
levels at doses of 200 mg/kg-day and greater. PFBS exposure did not alter maternal estradiol-,
progesterone-, or gonadotropin-releasing hormone. Reproductive hormone levels in males and
females were not evaluated by Lieder et al. (2009b). The changes in follicle and corpora lutea
development reported in the same study, however, may be associated with alterations in hormone
production/levels because ovarian follicles and corpora lutea produce estrogen and progesterone,
respectively (Foster and Gray. 2013; U.S. EPA. 1996a).
The hormonal effects observed in the NTP (2019) and Feng et al. (2017) studies might be
associated with adverse reproductive effects reported in these studies. Androgens, luteinizing
hormone, estradiol, and progesterone play an important role in normal development and in the
functioning of the female reproductive system (Woldemeskel. 2017; Foster and Gray. 2013).
Alterations in the levels and production of these reproductive hormones can disrupt endocrine
signals at the hypothalamic-pituitary level and lead to delayed reproductive development and
changes in functions (Rudmann and Foley. 2018; Woldemeskel. 2017; Foster and Gray. 2013).
Endpoint
Study
Exposure
Animal Group
Observation time

FII1S Reproductive Hormone Effects
Testosterone (T)
NTP 2018,4309741
28 Day Oral
Rat, Harlan Sprague-Dawley (9)
Day 28
• • •	a	•	*




Rat, Harlan Sprague-Dawley (cf)
Day 28
• • •


	«

Estrogen
Feng 2017, 3856465
20 Day Oral Gestation
P0 Mouse, ICR (9)
GD20
	


• Doses




Fl Mouse, ICR (9)
PNDI
	


/\. Significant Increase

PND30

V
T
Significant Descrease

PND60

~
—~
I—| Dose Range

Progesterone (P4)
Feng 2017, 3856465
20 Day Oral Gestation
P0 Mouse, ICR (9)
GD20
	

—~





Fl Mouse, ICR (9)
PNDI
• •	




PND 30
	

—~


PND6U

V
—~


Luteinizing Hormone (LH)
Feng 2017,3856465
20 Day Oral Gestation
FI Mouse. ICR (9)
PNDI
» • •	«

PND30
~A	A	~

PND60
M	1	•

Gonadotropin Releasing Hormone (GnRH)
Feng 2017, 3856465
20 Day Oral Gestation
PO Mouse, ICR (9)
GD20
	•	•




Fl Mouse, ICR (9)
PNDI
	•	•

PND30
	•	•

PND60
	•	•






)0 0 100
200
300 400 500 600
700 800 900 1,000 1,
00







Dose (mg/kg/day)

Figure 8. Reproductive Hormone Response to K+PFBS Exposure
(Click to see interactive data graphic and rationale for study evaluations for reproductive
hormone levels in HAWC)
39

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EPA/690/R-21/001F
4.2.2.4 Reproductive System Development, Including Markers of Sexual Differentiation and
Maturation (Female and Male)
Several measures of female reproductive development were affected by gestational
K PFBS exposure in mice (see Figure 9, Figure E-5, and Figure E-6). Feng et al. (2017)
reported a delayed first estrous in female PFBS-exposed offspring (>200 mg/kg-day) compared
with control (see Figure E-5). Estrous cyclicity was also affected in K+PFBS-exposed
PNDs 40-60 offspring as exhibited by a prolongation of the diestrus stage compared with
control. Estrous cycling was generally not statistically significantly altered in Po- or
Fi-generation females treated with K PFBS in the two-generation study by l.ieder et al. (2009b).
An increase in the number of rats with >6 consecutive days of diestrus was observed in the
Fi females exposed to 100 mg/kg-day; however, the increase was not present at higher doses
(l.ieder et al.. 2009b). Estrous cyclicity was affected after adult exposure to K PFBS for 28 days
as shown by a dose-dependent prolongation of diestrus at doses of 250 mg/kg-day and greater
with marginal significance at the lowest dose tested (125 mg/kg-day) (p = 0.063) (N I P. 2019).
l.ieder et al. (2009b) reported a delay in the days to preputial separation in Fi males of the
30- and 1,000-mg/kg-day groups;10 however, the measure was no longer statistically significant
when adjusted for BW. There was similarly no change in the days to vaginal patency in Fi
female rats (l.ieder et al.. 2009b). Unlike l.ieder et al. (2009b). Feng et al. (2017) reported a
delay in vaginal patency in Fi females after gestational exposure of 200 mg/kg-day and greater
(see Figure E-6).
10A marker of delayed reproductive development (Foster and Gray. 2013: U.S. EPA. 1996b).
40

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EPA 690 R-2 J 00 IF
Kndpoinl Name
Study Name
Experiment Name
Animal Description
Observation Time


PFBS Reproductive Development and F.strous Cycling


>6 Days of Diestrus
Liedcr. 2009, 1578545
2 Generation Oral
F1 Rat, Sprague-Dawley (9)



	•	
—4
• Doses




P0 Rat, Sprague-Dawley (9)
6 days



—~
|—j Dose Range

>6 Days of Estrus
Lieder. 2009. 1578545
2 Generation Oral
F1 Rat. Sprague-Dawley (9)


*~-

—4
/V Significant Increase




P0 Rat, Sprague-Dawley (9)
6 days



—~
Significant decrease

Bstrous Cycle, Diestrus
Feng 2017, 3856465
20 Day Oral Gestation
Fl Mouse, ICR (9)
PND40-60


	A	A



NTP 2018, 4309741
28 Day Oral
Ral, llarlan Sprague-Dawley 19)
Day 28

~—
—	A	A	
—~

Esirous Cycle. Esirus
Feng 2017, 3856465
20 Day Oral Gestation
Fl Mouse. ICR (9)
PND40-60


	•	~



NTP 2018, 4309741
28 Day Oral
Ral. Harlan Sprague-Dawley (9)
Day 28

~—

—4

Estrous Cycle, Metcstrus
NTP 2018. 4309741
28 Day Oral
Rat. Harlan Sprague-Dawley (9)
Day 28

M
• • •	
—~

Esirous Cycle. Procslrus
Feng 2017. 3856465
20 Day Oral Gestation
Fl Mouse, ICR (9)
PND40-60


	•	~



NTP 2018. 4309741
28 Day Oral
Rat. Harlan Sprague-Dawley (9)
Day 28

~—
• • •	
—~

Estrous Stage, Diestrus (at sacrifice)
Lieder, 2009, 1578545
2 Generation Oral
Fl Rat, Sprague-Dawley (9)
Sacrifice


-«	•	
—~




P0 Rat. Sprague-Dawley (9)


*~-
• ~	
—~

Estrous Stage, Esirus (at sacrifice)
Lieder, 2009, 1578545
2 Generation Oral
Fl Rat, Sprague-Dawley (9)
Sacrifice

*~-
• •	
—~




P0 Rat, Sprague-Dawley (9)


*•-
• •


Estrous Stage, Meieslrus (at sacrifice)
Liedcr. 2009. 1578545
2 Generation Oral
Fl Ral. Sprague-Dawley (9)
Sacrifice


-•	•
—•




P0 Rat, Sprague-Dawley (9)


M-
• •	
—4

Esirous Stage. Procslnis (al sacrifice)
Lieder. 2009. 1578545
2 Gcneralion Oral
Fl Ral. Sprague-Dawley (9)
Sacrifice


-m	•	
—~




IX) Rat, Sprague-Dawley (9)



-•	•	
—~

Esirous Stages/ 21 Days
Lieder. 2009, 1578545
2 Generation Oral
Fl Rat, Sprague-Dawley (9)


*~-
-«	•	
—~




P0 Rat, Sprague-Dawley (9)



-•	•	
—~

First Estrous - Litter N
Feng 2017, 3856465
20 Day Oral Gestation
Fl Mouse, ICR (9)
Beginning on PND24

~-*
	A	A


Preputial Separation
Lieder, 2009, 1578545
2 Generation Oral
Fl Rat, Sprague-Dawley ((f )
Day 29 (postpartum)


• •
-A

Vaginal Opening
Feng 2017. 3856465
20 Day Oral Gestation
Fl Mouse. ICR (9)
PND60

M	A	A

Vaginal Patency
Lieder, 2009. 1578545
2 Generation Oral
Fl Rat. Sprague-Dawley (9)
Postpartum Day 28

»• •	•	+






-100 0
100 200 300 4(H) 500 600 700 800 9(H)
1,000 1,1
00








Dose (mg/kg/day)


Figure 9. Effects on Reproductive Development and Estrous Cycling Following PFBS Exposure
(Click to see interactive data graphic)
41

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EPA/690/R-21/001F
4.2.2.5 Reproductive Organ Weights and Histopathology (Female and Male)
Studies have not consistently reported changes in reproductive organ weights (click to see
interactive graphic for reproductive organ effects in HAWC). Reproductive organ weights,
including testes, ovaries, and uterus, were unchanged in the two-generation reproductive study in
Po and Fi males and females (l.icdcr et al.. 2009b) and following short-term and sub chronic
exposure to K PFBS (N I P. 2019; l.ieder ct al.. 2009a; 3M. 2001. 2000d). Fi females
gestationally exposed to PFBS, however, exhibited decreased size and weight of the ovaries and
uterus (Feng ct al.. 2017). In addition, the total uterine section diameter and endometrial and
myometrial thickness were significantly reduced. There were no significant histopathological
alterations in the male or female reproductive organs evaluated following exposure to K+PFBS
for 28 days (N I P. 2019) or in parental or offspring from the two-generation reproductive study
(l.ieder ct al.. 2009b).
4.3 OFFSPRING GROWTH AND EARLY DEVELOPMENT
4.3.1 Human Studies
No human studies were available to inform the potential for PFBS exposure to cause
effects on the growth or early development of children.
4.3.2 Animal Studies
Evidence to inform organ/system-specific effects of PFBS in animals following
developmental exposure are discussed in the individual hazard sections (e.g., reproductive
cycling after developmental exposure is discussed in the "Reproductive Effects" section). This
section is limited to discussion of other, specific developmental effects commonly evaluated in
guideline developmental toxicity studies, including pup BW, developmental markers, and bone
measures. Four high- or medium-confidence studies examined potential alterations in offspring
growth and early development following PFBS exposure, including two gestational exposure
studies in rats (York. 2003a. 2002) and one gestational exposure study in mice (Feng et al..
2017). as well as a two-generation study in rats (l.ieder et al.. 2009b; York. 2003c). (Click to see
interactive graphic for developmental effects in HAWC.)
None of the studies identified significant effects in either rats or mice on measures of
fetal morphology (i.e., malformations and variations). BW of female offspring of PFBS-exposed
mice at doses greater than 200 mg/kg-day was statistically significantly lower than control at
PND 1, and the pups remained underweight through weaning, pubertal, and adult periods, with
decreases of approximately 25% observable in pups nearing weaning (Feng et al .. 2017). At
around PND 16, Feng et al. (2017) also reported an -1.5-day developmental delay in eye
opening in pups gestationally exposed to 200 mg/kg-day PFBS and greater. Importantly, no
effects on the health of the exposed dams (e.g., weight gain) were observed at any dose (Feng et
al.. 2017). Dose-response graphics for eye opening, including effect size and variability, are
included in Appendix E, Figure E-4. Fetal BWs (male and female) were also reduced
(approximately 10%) compared with controls following gestational exposure from GDs 6 to 20
at the highest tested dose (1,000 mg/kg-day in York (2002) and 2,000 mg/kg-day in York
(2003a)1). Parental BWs and organ weights, however, were also affected to a similar degree at
those doses (l.ieder et al.. 2009b; York. 2003c. 2002). limiting the interpretation of the results.
No statistically significant changes in Fi- and F2-generation pup mean pup weight at birth and
mean pup weight at weaning were reported by l.ieder et al. (2009b) or York (2003c).
42

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EPA/690/R-21/001F
Several measures of thyroid hormone development and female reproductive development
were affected by gestational PFBS exposure in mice and are described in more detail in the
"Thyroid Effects" and "Reproductive Effects" sections, respectively.
4.4 RENAL EFFECTS
4.4.1 Human Studies
One low-confidence study (Oin et al. (2016). with additional details in Bao et al. (2014).
selected 225 subjects ages 12-15 years old from a prior cohort study population in seven public
schools in northern Taiwan (Tsai et al.. 2010) and examined the association between PFBS
exposure and uric acid concentrations. There was no association between ln(PFBS)
concentration and uric acid concentrations in the total population (P: 0.0064 mg/dL increase in
uric acid per 1 ln-|ig/L increase in PFBS; 95% CI: -0.22-0.23). U.S. EPA identified that a
nonsignificant positive association in boys was offset by a nonsignificant negative association in
girls, and there is not enough information to determine whether there is a sex dependence. When
PFBS exposure was analyzed for high uric acid (>6 mg/dL), the risk was somewhat elevated in
boys (OR: 1.53; 95% CI: 0.92-2.54), but not in girls (OR: 0.99; 95% CI: 0.58-1.73). The
potential for reverse causation (i.e., that renal function could influence the levels of PFBS in the
blood) tempers any conclusions that might be drawn.
4.4.2 Animal Studies
Renal effects were evaluated in high-confidence short-term and subchronic-duration
exposure studies in rats (NTP. 2019: Lieder et al.. 2009a: 3M. 2001. 2000d) and in a
high-confidence two-generation reproductive study in rats (Lieder et al.. 2009b). Endpoints
evaluated in these studies include kidney weights, histopathological changes, and serum
biomarkers of effect (see Figure E-8 and Figure E-9). Dose-response graphics for
histopathological effects, including effect size and variability, are included in Appendix E,
Figure E-7.
Absolute and relative kidney weights of males and females were unchanged in S-D rats
exposed daily for 90 days to K+PFBS at doses up to 600 mg/kg-day compared with control rats
(Lieder et al .. 2009a). This lack of effect on kidney weight was also observed in parental and Fi
male and female rats of the same strain exposed to K+PFBS at doses up to 1,000 mg/kg-day
during a two-generation reproductive study (l.ieder et al.. 2009b). Although none of the findings
reached statistical significance, an approximate 9% increase in absolute kidney weight was
observed in female S-D rats exposed to 1,000 mg/kg-day K PFBS for 10 days (3M. 2000d);
relative-to-body kidney weights were also increased approximately 6—9%. This organ-weight
effect was not observed in corresponding males of the study. In a follow-up 28-day study by the
same lab, a 9-11% increase in absolute and relative-to-body kidney weight was observed in
female S-D rats exposed to 900 mg/kg-day K PFBS (3M. 2001). although these changes were
not statistically significant. In this study, U.S. EPA also observed that smaller nonsignificant
increases in kidney weight occurred in male rats. In another 28-day study, K+PFBS exposure
significantly increased absolute and relative right kidney weights in high-dose (500 mg/kg-day)
male S-D rats (NTP. 2019). Only relative kidney weights were altered in female rats, but this
effect was significant at all tested K+PFBS doses (>62.6 mg/kg-day). Click to see interactive
graphic for kidney-weight effects in HAWC.
43

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EPA/690/R-21/001F
After 90 days of exposure, Lieder et al. (2009a) observed increased incidences of
histopathological alterations of the kidneys of male and female rats of the high-dose group
(600 mg/kg-day). Increased incidence of hyperplasia of the epithelium of renal papillary tubules
and ducts was observed in rats of both sexes (see Figure E-7Figure E-8). A single incidence of
papillary necrosis in both kidneys was observed in one male in the high-dose group. Further,
focal papillary edema was observed in 3/10 rats of both sexes of the high-dose groups compared
with no evidence of this effect in control rats. Similar histopathological alterations were
observed in parental and Fi male and female rats in the two-generation reproduction study
(l.ieder et al.. 2009b). Compared with control rats, increased incidences of hyperplasia of the
renal tubular and ductal papillary epithelium, and focal papillary edema were observed in
parental male and female rats at PFBS doses >300 mg/kg-day. Hyperplastic foci in the same
locations of the kidney were also observed in male and female Fi rats exposed to
>300 mg/kg-day PFBS across life stages from gestation to adulthood (l.ieder et al.. 2009b).
Focal papillary edema was observed in male (>1,000 mg/kg-day) and female (>300 mg/kg-day)
Fi rats, although this specific alteration did not appear to be dose-dependent in females.
Although kidney alterations such as hydronephrosis, mineralization, and tubular degeneration
were observed in male or female S-D rats after just 10 days of oral K+PFBS exposure, these
effects were not significant compared to control and/or did not appear to be dose-dependent (3M.
2000d). The same histopathological lesions were noted in the 28-day rat study albeit with lack
of statistical significance compared to control (3M. 2001). In another 28-day gavage study in
S-D rats, chronic progressive nephropathy (CPN) was observed in all male and female PFBS
treatment groups and control rats, with no evidence of dose dependence for this effect (N I P.
2019). Renal papillary necrosis was also observed in these rats but only at the highest exposure
dose (1,000 mg/kg-day).
Serum levels of biomarkers indicative of kidney injury and/or function, including blood
urea nitrogen (BUN) and creatinine, have been examined across multiple studies of varying
exposure durations, and were found to be unchanged in male and female rats treated with
K PFBS at doses up to 1,000 mg/kg-day (l.ieder et al.. 2009a; 3M. 2001. 2000d). After 28 days
of gavage exposure in S-D rats, however, N I P (2019) observed significantly increased levels of
BUN in males (>250 mg/kg-day). This increased circulating BUN was not observed in female
rats at doses up to 1,000 mg/kg-day. Click to see interactive graphic for other kidney effects in
HAWC.
4.5 HEPATIC EFFECTS
4.5.1 Human Studies
No human studies were available to inform the potential for PFBS exposure to cause
hepatic effects.
4.5.2 Animal Studies
Hepatic effects were evaluated in high-confidence short-term and subchronic studies in
rats (N I P. 2019; l.ieder et al.. 2009a; 3M. 2001. 2000d) and in a high-confidence two-generation
reproductive study in rats (l.ieder et al.. 2009b). Endpoints evaluated in these studies include
liver weights, histopathological changes, and serum biomarkers of effect (see Figure E-10).
44

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EPA/690/R-21/001F
Ten days of daily gavage exposure to K+PFBS significantly increased absolute,
relative-to-body, and relative-to-brain weights of liver in adult male and female S-D rats exposed
to 1,000 mg/kg-day (3M, 2000d). The absolute liver mass of male rats was increased by 36%
compared with females (22%). A similar profile of liver-weight alteration in S-D rats was
observed following 28 days of exposure wherein absolute and relative liver weights of high-dose
(900 mg/kg-day) male rats had increased 25%—30% (3M, 2001). Female rats at the same
treatment dose did not experience a similar magnitude increase in absolute or relative liver
weights (4~6%). In another 28-day study in S-D rats, K+PFBS exposure significantly increased
absolute and relative liver weights in males (>125 and >62.6 mg/kg-day, respectively) and
females (>250 and >125 mg/kg-day, respectively) (NIP. 2019). In contrast, the livers of male
and female S-D rats exposed to K+PFBS at doses up to 600 mg/kg-day for 90 days were not
significantly changed compared with respective controls (l.ieder et al.. 2009a). In a
two-generation reproduction study using the same strain of rat, however, increased absolute and
relative liver weights were observed in male parental rats exposed to doses of K+PFBS
>300 mg/kg-day for approximately 70 days (l.ieder et al.. 2009b). In the Fi adult males, only
relative liver weight was significantly increased at the high dose (1,000 mg/kg-day), although
terminal BW was significantly decreased in this group compared with control.
Histopathological examination of the livers of S-D rats across three separate gavage
studies of increasing K PFBS exposure duration [ 10-day, 3M (2000d); 28-day, 3M (2001);
90-day, l.ieder et al. (2009a) 1 did not reveal any significant dose-dependent alterations or lesions.
For example, focal/multifocal hepatic inflammation was observed in 3/10 male and 4/10 female
rats of the high-dose group (no incidence at the low or mid dose) compared to 6/10 male and
female rats in the control groups (l.ieder et al.. 2009a). The l.ieder et al. (2009b) two-generation
reproduction gavage study did identify increased incidences of hepatocellular hypertrophy in
parental and Fi adult male rats at >300 mg/kg-day; however, this effect was absent in female rats
at doses of K PFBS up to 1,000 mg/kg-day. N I P (2019) identified a significantly increased
incidence of hepatocellular hypertrophy in male (>125 mg/kg-day) and female (>500 mg/kg-day)
S-D rats after 28 days of K+PFBS exposure. Further, significantly increased cytoplasmic
alteration of hepatocytes was observed in these rats (male and female at >500 mg/kg-day).
Hepatic necrosis was also observed but was not significant compared with control and only
occurred at the high dose (1,000 mg/kg-day) in both sexes (N I P. 2019).
In general, serum biomarkers associated with altered liver function or injury, including
alanine aminotransferase (ALT) and aspartate aminotransferase (AST), were not significantly
changed in male and female S-D rats across multiple gavage studies of varying exposure
durations up to 90 days and at K PFBS doses up to 1,000 mg/kg-day (l.ieder et al.. 2009a; 3M.
2001. 2000d). N I P (2019). however, reported increased serum ALT and AST in male
(500 mg/kg-day only) and female (>250 mg/kg-day for ALT; >500 mg/kg-day for AST) rats
exposed to K PFBS for 28 days. Click to see interactive graphic for liver effects in HAWC.
4.6 EFFECTS ON LIPIDS OR LIPOPROTEINS
4.6.1 Human Studies
One low-confidence study (Zeng et al.. 2015) used the controls from the case-control
study of asthma described below (Dong et al .. 2013a) and examined the association between
PFBS exposure and serum lipids. There was a statistically significant increase in total
45

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EPA/690/R-21/001F
cholesterol (P: 19.3 mg/dL increase per 1 |ig/L increase in PFBS; 95% CI: 0.6-38.0) but when
PFBS exposure was analyzed in quartiles, no exposure-response gradient was observed.
In addition, a medium-confidence birth cohort study in China examined associations with
childhood adiposity (Chen et al.. 2019). PFBS was measured in cord blood samples at birth and
several measures of adiposity were collected at age 5 years. There was higher adiposity with
higher exposure in girls, with significant exposure-response relationships across tertiles with
waist circumference, fat mass, body fat percentage, and waist-to-height ratio. No association
with adiposity was observed in boys. It is unlikely that the association in girls can be explained
by confounding across the other PFAS measured in this study as the associations were strongest
for PFBS, but it is possible that there is other unmeasured confounding.
4.6.2 Animal Studies
Beyond a single medium-confidence mouse study fBiiland et al. (2011); 3M (2010);
summarized below], PFBS studies have not particularly focused on perturbations in lipids or
lipoproteins as a potential health outcome, because studies have typically focused only on
measures of serum cholesterol and triglyceride as part of a broader panel of clinical chemistry
measures in high- or medium-confidence rat studies of 10, 28, and 90 days (see Figure E-ll)
r3M (2000d). 3M (2001). and l.ieder et al. (2009a). respectively]. Circulating levels of
cholesterol and triglycerides were unchanged in male and female S-D rats following daily
gavage exposure to K PFBS for 10 days at doses up to 1,000 mg/kg-day (3M. 2000d). In a
similarly designed study from the same laboratory, serum cholesterol and triglyceride levels
were decreased in male rats but at the high dose only, and this effect was neither statistically
significant compared with control nor observed in female rats of the same dose group (3M.
2001). Following exposure for up to 90 days, cholesterol and triglycerides were unchanged in
male and female rats at doses up to 600 mg/kg-day (l.ieder et al.. 2009a). PFBS was included in
a multi-PFAS study specifically designed to interrogate the mechanism of effect on lipid and
lipoprotein metabolism in a transgenic mouse line (APOE*3-Leiden CETP) that is highly
responsive to fat and cholesterol intake, consistent with human populations exposed to a
western-type diet (containing 14% beef tallow, 1% corn oil, and 0.25% cholesterol) (Bijland et
al.. 2011; 3M. 2010). Adult male mice were fed a western-type, high-fat diet for 4 weeks prior
to initiation of PFBS exposure and throughout the 4- to 6-week PFBS exposure period (at
approximately 30 mg/kg-day). This study included several measures of lipid and lipoprotein
synthesis, modification, and transport or clearance, such as circulating plasma levels, in vivo
clearance of very low-density lipoprotein (VLDL)-like particles, fecal bile acid and sterol
excretion, hepatic lipid levels, lipase activity, VLDL-triglyceride and VLDL-apoB production,
and gene expression profiles. After 4 weeks of PFBS exposure, fasting plasma triglycerides,
cholesteryl ester transfer protein, and glycerol were significantly decreased compared with mice
on the control diet. Further, the half-life of VLDL-like particles and hepatic lipase activity, and
hepatic cholesteryl ester and free cholesterol levels were decreased (Bijland et al.. 2011; 3M.
2010). Hepatic uptake of VLDL-like particles (represents fatty acid/lipid transport into hepatic
tissue) was modestly, but significantly, increased compared with control mice. This increased
hepatic lipid uptake in the liver was accompanied by increased expression of genes associated
with lipid binding, activation, and metabolism (e.g., P-oxidation).
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4.7 OTHER EFFECTS
4.7.1 Human Studies
Two studies in China examined different immune outcomes in children (Chen et al..
2018; Dong et al.. 2013a).
One medium-confidence study reported in five publications (Oin et al.. 2017; Zhou et al..
2017b; Zhou et al.. 2017a; Zhu et al.. 2016; Dong et al.. 2013b) examined the association
between PFBS exposure and asthma, asthma symptoms, pulmonary function, and related
immune markers (immunoglobulin E [IgE], absolute eosinophil count [AEC], eosinophilic
cationic protein [ECP], T-helper cell-specific cytokines, and 16-kDa club cell secretory protein).
The primary finding was a statistically significant (in the fourth quartile) positive association
between incident asthma (i.e., diagnosis in the previous year) and PFBS exposure (OR for Q2:
1.3, 95% CI: 0.7-2.3; ORforQ3: 1.2, 95% CI: 0.7-2.2; ORforQ4: 1.9, 95% CI: 1.1-3.4).
There were also increases in AEC and ECP with increased exposure (not statistically significant
with the exception of AEC in children with asthma). There was no clear association with IgE or
T-helper cell-specific cytokines. There was also no clear association with asthma severity or
control of asthma symptoms (Dong et al.. 2013a). or pulmonary function measured with
spirometry among children with asthma (Oin et al.. 2017). While reduced pulmonary function
could be considered an outcome separate from asthma, the study authors noted no associations in
pulmonary function (i.e., in nonasthmatics across the PFAS they studied), so for these purposes,
it was considered an indicator of asthma severity.
One medium-confidence study (Chen et al.. 2018) examined the association between
PFBS exposure and atopic dermatitis and reported a statistically nonsignificant increase in atopic
dermatitis with increased exposure (OR: 1.23; 95% CI: 0.74-2.04).
In addition, two studies examined cardiovascular effects (Huang et al.. 2019b; Huang et
al.. 2018). but it is difficult to evaluate consistency across studies given the different outcomes in
each.
One medium-confidence study (Huang et al.. 2018) using data from NHANES cycles for
1999-2014 reported significantly higher odds of total cardiovascular disease with higher
exposure (OR for above vs. below the LOD: 1.19; 95% CI: 1.06-1.32) and elevated, though not
statistically significant, odds of individual types of cardiovascular disease (congestive heart
failure, coronary heart disease, angina pectoris, heart attack, and stroke). There is potential in
this study for confounding across the PFAS, because PFBS was highly correlated with some
other PFAS with slightly stronger associations.
A medium-confidence cross-sectional study (Huang et al.. 2019b) of hypertensive
disorders of pregnancy reported higher odds for all such disorders in pregnancy (in the third
tertile) (OR for Tertile 2 vs. Tertile 1: 0.89, 95% CI: 0.39-2.44; OR for Tertile 3: 2.26, 95% CI:
1.02-5.0; p-trend 0.03) and pre-eclampsia (OR for Tertile 2 vs. Tertile 1: 2.09, 95% CI:
0.51-8.53; OR for Tertile 3: 3.51, 95% CI: 0.94-13.2;/Mxend 0.05), with both trends being
statistically significant after mutual adjustment of PFAS.
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4.7.2 Animal Studies
Other effects were evaluated following exposure to PFBS, including outcomes related to
the spleen, hematological system, BW, neurotoxicity, and nonspecific clinical chemistry. These
groups of outcomes were not synthesized because of inadequate available information, uncertain
biological relevance, and/or inconsistencies across studies and sexes.
4.8 OTHER DATA
Other studies that used PFBS or K+PFBS are described in this section. These studies are
not adequate for determining RfD values and were considered supportive data. These data might
include acute-duration exposures, genotoxicity, mechanistic, and other studies (see Table 5).
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Table 5. Other Studies
Test
Materials and Methods
Results
Conclusions
References
Genotoxicity
Mutagenicity
test
Salmonella typhimurium (strains TA98 and TA100) and
Escherichia coli (strain pKMlOl) in the presence or
absence of S9. Concentrations of PFBS were between
0-5,000 (ig/plate.
Test was negative for TA100 and pKMlOl
strains and equivocal for TA98 strain.
There is no in vitro
evidence of PFBS
mutagenicity.
NTP (2005)
Ames
S. typhimurium (strains TA98, TA100, TA1535, and
TA1537) and E. coli (strain WP2uvrA) were tested in
the presence or absence of S9 and with or without a
preincubation treatment. Concentrations of K+PFB S
were between 0-5,000 (ig/plate.
The results of both mutation assays
indicate that PFBS did not induce any
significant increase in the number of
revertant colonies for any of the tester
strains in the presence or absence of
induced rat liver S9.
There is no in vitro
evidence of PFBS
mutagenicity.
Pant (2001)
Genotoxicity
test
Human hepatoma (HepG2) cells were treated with
0.4 |iIVI to 2 mM PFBS. Intracellular ROS production
was measured by use of 2',7'-dichlorofluorescein
diacetate and DNA damage was measured with the
comet assay.
The amount of ROS and DNA strand
breaks remained unaffected by PFBS
treatment.
PFBS did not generate
ROS or DNA damage in
human liver cells.
Erikseti et al.
(2010)
CHO
chromosomal
aberration
Cultures of CHO cells were treated with K+PFBS at
concentrations ranging from 0 to 5,000 (ig/mL with or
without exogenous metabolic activation. The in vitro
exposure duration was 3 hr.
PFBS did not induce a statistically
significant increase in the percentage of
cells with aberrations at any of the
concentrations tested, either with or
without metabolic activation, in either
assay when compared to the solvent
controls.
Based on the negative
results in the in vitro CA
assay in CHO cells, PFBS
is not considered to be a
clastogenic agent.
Xu (2001)

Micronucleus
assay
Male and female S-D rats (5/group) were exposed twice
daily to K+PFBS by gavage at doses of 31.3, 62.5, 125,
or 250 mg/kg for 28 d.
PFBS did not induce a statistically
significant increase in the frequency of
micronucleated polychromatic
erythrocytes.
PFBS was negative for
micronuclei in the blood
of male and female rats,
indicating a lack of
genotoxic potential.
NTP (2012)
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Table 5. Other Studies
Test
Materials and Methods
Results
Conclusions
References
Acute duration and other routes of exposure
Acute
10 rats/group, young adult male rat (strain not
specified), administered PFBS by gavage, single dose,
50, 100, 300, 600, or 800 (iL/kg and observed for 14 d
postexposure.
Mortality: 0, 20, 60, 80, and 100% at 50,
100, 300, 600, and 800 pL/kg PFBS,
respectively.
Acute oral PFBS rat LD5o
in male rats is 236 (iL/kg
(corresponding to
430 mg/kg).
Bombard and
Loser (1996)
Low
confidence
Acute dermal
Adult (8 wk of age) male and female S-D rats (5/group)
were exposed dermally (10% of body surface area) to
500, 1,000, or 2,000 mg/kg K+PFBS for 24 hr and then
observed for 15 d postexposure for signs of clinical
toxicity, mortality, BW changes, or gross pathology
(terminus of study).
No treatment-related observations were
noted.
PFBS is not acutely toxic
via the dermal route of
exposure in rats.
3M (2000b)
Dermal irritation
Adult (14-wk of age) female NZW rabbits (3 rabbits
total for study) were exposed dermally (6 cm2 of skin) to
500 mg K+PFBS for approximately 4 hr and then
observed for 9 d postexposure for signs of clinical
toxicity, mortality, or BW changes.
Draize scoring was performed on the patch
site immediately following the exposure
period and 24, 48, and 72 hr postexposure.
No signs of dermal irritation were
observed. No signs of clinical toxicity or
mortality occurred. No treatment-related
alterations in B W were noted.
PFBS did not induce
erythema, edema, or other
possible dermal findings
during the scoring periods,
indicating a lack of dermal
irritant properties in
rabbits.
3M (2000a)
Ocular
sensitivity
Adult (16-wk of age) female NZW rabbits (3 rabbits
total for study) were exposed to approximately 80 mg
K+PFBS via ocular installation in the left eye for 2 sec.
Eyes were flushed with 0.9% saline after 24 hr and then
observed and scored for up to 21 d postexposure. The
rabbits were also followed for clinical signs of toxicity
or mortality/moribundity.
Excessive lacrimation of the left eyes
noted throughout study postexposure.
Based on the laboratory scoring system,
PFBS was "moderately" irritating at 24
and 72 hr postexposure.
PFBS is a moderate ocular
irritant in rabbits.
3M (2000c)
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Table 5. Other Studies
Test
Materials and Methods
Results
Conclusions
References
Contact
hypersensitivity
Adult male (10-12 wk old) and female (9 wk old)
CRL:(HA)BR Hartley guinea pigs were injected
intradermally with sterile water, Freund's adjuvant, or
adjuvant containing 125 mg/mL K+PFBS (induction
phase). D 7 after induction, a petrolatum paste containing
0.5 g K+PFBS was applied to the previous injection site of
the guinea pigs for 48 hr (topical induction phase). D 22,
a challenge dose of 0.5 g K+PFBS (petrolatum paste) was
applied to the shaved left cranial flank (right flanks were
treated with petrolatum paste only) (challenge phase).
This challenge procedure was repeated on D 29.
Challenge sites were observed and scored following each
challenge period (D 24-25 males and females
and D 31-32 males only). Guinea pigs were also
followed for signs of clinical toxicity,
mortality/moribundity, or alterations in BW.
No mortalities, clinical signs of toxicity, or
changes in BW associated with PFBS
exposure were noted. Dermal scores were
zero (no response) in females and did not
exceed 1 in males (discreet or patchy
edema), which was not considered
significant compared with control guinea
pigs exposed to Freund's adjuvant alone.
PFBS is not considered an
allergen in the guinea pig
maximization test.
3M (2002a)
BW = body weight; CA = chromosomal aberration; CHO = Chinese hamster ovary; DNA = deoxyribonucleic acid; K+PFBS = potassium perfluorobutane sulfonate;
LD5o = median lethal dose; NZW = New Zealand White; PFBS = perfluorobutane sulfonic acid; ROS = reactive oxygen species; S-D = Sprague-Dawley.
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4.8.1	Tests Evaluating Genotoxicity and Mutagenicity
Genotoxic, mutagenic, and clastogenic effects of PFBS have been tested in mammalian
and prokaryotic cells in vitro (Hriksen et al.. 2010; N I P. 2005; Pant 2001; Xu. 2001). and in rats
in vivo (N I P. 2019). PFBS was negative for mutagenicity in Escherichia coli strain pKM 101
and Salmonella typhimurium strain TA100 (N'l'P. 2005). Mutagenicity test results were
equivocal in S. typhimurium strain TA98. Pant (2001) tested PFBS at concentrations up to
5,000 [j,g/plate in E. coli strain WP2uvrA and S. typhimurium strains TA98, TA100, TA1535,
and TA1537 in the presence or absence of exogenous metabolic activation and found no
evidence of mutagenic activity. In mammalian cells in vitro, PFBS did not generate reactive
oxygen species (ROS) or oxidative deoxyribonucleic acid damage in HepG2 cells (Erikscn et al..
2010). PFBS also failed to induce chromosomal aberrations in Chinese hamster ovary cells,
suggesting a lack of clastogenic activity (Xu. 2001). Adult male and female S-D rats exposed
twice daily to oral PFBS at doses up to 250 mg/kg for 28 days did not experience any significant
increases in micronucleated polychromatic erythrocytes, indicating a lack of genotoxic activity
(see Table 5) (N I P. 2012).
4.8.2	Acute Duration and Other Routes of Exposure
Limited data are available to evaluate acute toxicity and effects from dermal exposure to
PFBS (see Table 5). One low-confidence acute oral toxicity study on male rats administered
PFBS by gavage reported a median lethal dose (LDso) of 236 (iL/kg (corresponding to
430 mg/kg) (Bomhard and Loser. 1996). One acute dermal toxicity study concluded that PFBS
is not acutely toxic via the dermal route of exposure in rats, with no treatment-related
observation at doses up to 2,000 mg/kg (3M. 2000b). PFBS was not reported to induce
erythema, edema, or other possible dermal findings during the scoring periods, indicating a lack
of dermal irritant properties in rabbits exposed to 500 mg K+PFBS for approximately 4 hours
(3M, 2000a). PFBS was found to be a moderate ocular irritant in rabbits exposed to 80 mg
K PFBS via ocular installation (3M. 2000c). PFBS did not induce skin sensitization in the
guinea pig maximization test with an intradermal injection of 125 mg/mL and topical induction
of 0.5 g K PFBS (3M. 2002a).
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5.0 EVIDENCE INTEGRATION AND HAZARD CHARACTERIZATION
The epidemiology database of studies of PFBS exposure and health effects consists of
19 epidemiologic studies (described in 22 publications), summarized in the previous section.
The experimental animal database of all repeated-dose oral toxicity studies for PFBS and the
related compound K PFBS includes a short-term range-finding study in rats (3M, 2000d). two
28-day studies in rats (N I P. 2019; 3M, 2001). one subchronic study in rats (l.ieder et al.. 2009a).
one subchronic-duration lipoprotein metabolism study in mice (Bijland et al.. 2011; 3M. 2010).
three gestational exposure studies in mice and rats (Feng et al.. 2017; York. 2003a. 2002). and a
two-generation reproductive toxicity study in rats (l.ieder et al.. 2009b). Health outcomes
evaluated across available studies included effects on the thyroid, reproductive organs and
tissues, developing offspring, kidneys, liver, and lipids/lipoproteins following oral exposure to
PFBS. Table 6 provides an overview of this database of potentially relevant studies and effects.
This table includes only the high- and medium-confidence animal studies (a single,
low-confidence animal study was not considered informative for drawing conclusions on
potential health hazard[s]). The available epidemiology studies are also not included because
their ability to inform conclusions about associations was limited because of the small number of
studies (typically one) per outcome and poor sensitivity resulting from low exposure levels.
Following the summary of the available database in Table 6, narrative summaries
describe the evidence integration judgments and the primary rationales supporting these
decisions for each health effect. These narratives are supported by an evidence profile table that
succinctly lays out the various factors that were judged to increase or decrease the support for a
hazard. While the epidemiology studies were not influential in drawing evidence integration
judgments (i.e., they were judged as equivocal for all outcomes) or the derivation of toxicity
values (i.e., these studies are not discussed in the next section), the general findings are
summarized below to provide context to the animal study findings and identify potential areas of
future research.
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Table 6. Summary of Noncancer Data for Oral Exposure to PFBS (CASRN 375-73-5) and the Related Compound
K+PFBS (CASRN 29420-49-3)
Exposure
Duration3
Reference
Study
Confidence
Number of
Male/Female, Strain,
Species, Study Type,
Study Duration
Doses
Tested
(mg/kg-d)
Effects Observed at LOAEL
NOAEL
(mg/kg-d)
LOAEL
(mg/kg-d)
Short term
3M (2000d)
Medium
confidence
5/5, S-D rat, K+PFBS
administered by
gavage, 10 d
0, 100, 300,
1,000
Increased absolute and relative liver weight.
300
1,000
Short term
3M (2001)
High
confidence
10/10, S-D rat,
K+PFBS administered
by gavage, 28 d
0, 100, 300,
900
Increased absolute and relative liver weight
(male) and relative kidney weight (female).
300
900
Short term
NTP (2019)
High
confidence
10/10, S-D rat, PFBS
administered by
gavage, twice/d, 28 d
0, 62.6, 125,
250, 500,
l,000b
Decreased T3, free T4, total T4 in males and
females. Increased relative liver weight in
females and increased relative right kidney
weight in males.
NDr
62.6
Subchronic
Lieder et al.
(2009a): York
(2003b)
High
confidence
10/10, S-D rat,
K+PFBS administered
by gavage, 7 d/wk,
90 d
0, 60, 200,
600
Increased incidence of renal hyperplasia in
males and females.
200
600
Subchronic
Biiland et al.
(2011); 3M
(2010)
Medium
confidence
6-8/0, Apoe*3-Leiden
CETP mice, K+PFBS
in diet, 4-6 wk
0,30
Alterations in lipid homeostasis (e.g., decreased
hepatic lipase, triglycerides) is of uncertain
biological significance.
NDr
NDr
Developmental
Feng et al.
(2017)
High
confidence
0/10, ICR mice,
K+PFBS administered
by gavage, GDs 1-20
0, 50, 200,
500
Decreased T3, free T4, and total T4 in dams and
PND 1, 30, and 60 offspring. Increased TSH in
maternal and offspring (PND 30 only).
Delayed eyes opening, vaginal opening, and
first estrous and decreased BW in pups.
50
200
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Table 6. Summary of Noncancer Data for Oral Exposure to PFBS (CASRN 375-73-5) and the Related Compound
K+PFBS (CASRN 29420-49-3)
Exposure
Duration3
Reference
Study
Confidence
Number of
Male/Female, Strain,
Species, Study Type,
Study Duration
Doses
Tested
(mg/kg-d)
Effects Observed at LOAEL
NOAEL
(mg/kg-d)
LOAEL
(mg/kg-d)
Developmental
York (2003a)
High
confidence
0/8, S-D rat, K+PFBS
administered by
gavage, GDs 6-20
0, 100, 300,
1,000, 2,000
Decreased maternal feed consumption, BW
gain, and gravid uterine weight. Decreased pup
BW occurred at doses affecting maternal
health, limiting the interpretation of the results;
thus, developmental effect levels were not
determined. (Limited endpoints
evaluated—pilot study.)
P0: 1,000
Fi: NDr
P0: 2,000
Fi: NDr
Developmental
York (2002)
High
confidence
0/25, S-D rat, K+PFBS
administered by
gavage, GDs 6-20
0, 100, 300,
1,000
Decreased maternal feed consumption and BW
gain. Decreased pup BW occurred at doses
affecting maternal health, limiting the
interpretation of the results; thus,
developmental effect levels were not
determined.
P0: 300
Fi: NDr
P0: 1,000
Fi: NDr
Reproductive
Lieder et al.
(2009b): York
(2003c): York
(2003d): York
(2003e)
High
confidence
30/30, S-D rat,
K+PFBS administered
by gavage,
two-generation
reproductive study
P0 adults: 0,
30, 100,
300, 1,000
Fi adults: 0,
30, 100,
300, 1,000
P0 and Fi adults: increased incidence of
hyperplasia and focal papillary edema in the
kidneys of males and females.
F2 pups: no dose-related effects at the highest
dose tested (1,000 mg/kg-d).
Po, Fi: 100
F2: 1,000
Po, Fi:
300
F2: NDr
aDuration categories are defined as follows: Acute = exposure for <24 hours; short term = repeated exposure for 24 hours to <30 days; long term
(subchronic) = repeated exposure for >30 days <10% lifespan for humans (>30 days up to approximately 90 days in typically used laboratory animal species);
chronic = repeated exposure for >10% lifespan for humans (>~90 days to 2 years in typically used laboratory animal species) (U.S. EPA. 2002).
' Rats were gavaged twice daily at administered doses of 0, 31.3, 62.6, 125, 250, and 500 mg/kg in NTP (2019).
BW = body weight; GD = gestation day; NDr = not determined; ICR = Institute of Cancer Research; K+PFBS = potassium perfluorobutane sulfonate;
LOAEL = lowest-observed-adverse-effect level; NOAEL = no-observed-adverse-effect level; PFBS = perfluorobutane sulfonic acid; PND = postnatal day;
S-D = Sprague-Dawley; T3 = triiodothyronine; T4 = thyroxine; TSH = thyroid-stimulating hormone.
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5.1 THYROID EFFECTS
PFBS-induced perturbation of the thyroid was consistently observed across two species,
sexes, life stages, and exposure durations in two independent, high-confidence studies. These
perturbations involved a coherent pattern of hormonal changes. Significant changes in tissue
weight or histopathology were not observed.
Similar patterns of decreases in total T.<. total T-t. and free T-t were observed in
PFBS-exposed pregnant mice, nonpregnant adult female and adult male rats from a 28-day
study, and gestationally exposed female mouse offspring (N I P. 2019; Feng et al.. 2017). These
decreases were statistically significant (-20% in dams and -50% in offspring) and shown to
persist at least 60 days after gestational exposure in offspring and exhibited dose dependence in
both studies.
Development of numerous organ systems, including neuronal, reproductive, hepatic, and
immune systems, is affected by altered thyroid homeostasis because adequate levels of thyroid
hormones are necessary for normal growth and development in early life stages (Forhead and
Fowden. 2014; Gilbert and /ocller. 2010; Hulbert. 2000). Thus, the observed effects of PFBS
exposure on thyroid hormone economy are biologically consistent with the reported delays and
abnormalities in organ/system development discussed below. It is well established that the
presence of sufficient thyroid hormones during the gestational and neonatal period is essential
for brain development and maturation. Studies specifically evaluating the effect of PFBS on
neurodevelopment were not identified, leaving uncertainty as to the potential for adverse
developmental effects. Nonetheless, the coherence of these PFBS findings, in addition to the
large number of xenobiotic exposure studies demonstrating associations between thyroid
hormone economy and decrements in early life stage growth, development, and survival,
provides support for thyroid hazard.
Taken together, the evidence in animals for thyroid effects supports a hazard. The single
available study in humans did not report an association with thyroid hormones, but had severe
limitations hindering its interpretation. This low-confidence cross-sectional study was conducted
in a highly selected population (i.e., women with premature ovarian insufficiency), had poor
sensitivity, and methodological limitations (Zhang et al.. 2018). The limited evidence for thyroid
effects in human studies is equivocal. Although there are some differences in
hypothalamic-pituitary-thyroid (HPT) regulation across species (e.g., serum hormone-binding
proteins, hormone turnover rates, and timing of in utero thyroid development), rodents are
generally considered to be a good model for evaluating the potential for thyroid effects of
chemicals in humans (/ocller et al.. 2007). For more details pertaining to HPT dynamics and the
similarities and differences associated with thyroid hormone economy between rodents and
humans, please refer to A Literature Review of the Current State of the Science Regarding
Species Differences in the Control of and Response to, Thyroid Hormone Perturbations. Part 1:
A Human Health Perspective (Regulatory Science Associates. 2019). The pattern of decreased
thyroid hormones in the absence of a coordinated reflex increase in TSH and commensurate
alterations in thyroid tissue weight and/or histology, observed in PFBS studies [e.g., Feng et al.
(2017)1. is consistent with the human clinical condition referred to as "hypothyroxinemia,"
which is commonly associated with pregnancy in humans. Hypothyroxinemia has been defined
as a low percentile value of FT4 (ranging from the 2.5th percentile to the 10th percentile of FT4),
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with a TSH level within the normal reference range (Hales et al.. 2018; Alexander et al.. 2017;
Lazarus et al.. 2012; Negro et al.. 2011). Overall, based on findings in animal models considered
to be informative for evaluating the potential for thyroid effects in humans, the available
evidence supports a hazard, and the thyroid is considered a potential target organ for PFBS
toxicity in humans.
5.2 DEVELOPMENTAL EFFECTS
Overt effects on birth parameters and early development have generally not been
observed in either rats or mice after PFBS exposure. Specifically, the available studies do not
provide evidence of effects on endpoints relating to pregnancy loss, fetal survival, or fetal
morphology (Feng et al.. 2017; Lieder et al.. 2009a; York. 2003a. c, 2002). While one mouse
study indicated pronounced decreases in female offspring BW at several ages after gestational
exposure (Feng et al.. 2017). several other studies either did not observe decreases in offspring
BW or only detected these changes when parental BWs were similarly affected (Feng et al..
2017; l.ieder et al.. 2009a; York, 2003a. c, 2002).
Delays in development have been reported following gestational PFBS exposure in mice,
including delayed development of the female reproductive organs (i.e., ovaries, uterus, and
vaginal patency), delayed and abnormal estrous cycling (i.e., first estrous and prolongation of
diestrus), and delayed eye opening (Feng et al.. 2017). Age at vaginal patency and ovarian
follicle counts (i.e., in Fi rat offspring after delivery of the F2 generation) were unaffected at
1,000 mg/kg-day in a two-generation reproductive toxicity study (l.ieder et al.. 2009a). This
observed lack of effects (i.e., on vaginal patency) is inconsistent with the findings in mice.
However, Feng et al. (2017) also noted changes in reproductive hormones that might be relevant
to the delays in female sexual development, including a decrease in serum estradiol and
increased luteinizing hormone in pubertal offspring (i.e., PND 30 [Note: progesterone was
decreased at a later age, PND 60, but not PND 30]). Because the changes reported in mice by
Feng et al. (2017) were observed in parallel with effects on thyroid hormone levels (discussed
above), it is plausible that these developmental delays and hormonal changes could represent
sequalae of reduced thyroid function, although that was not directly tested.
For the most part, developmental effects have been reported in a single study and species
(mouse); however, the findings are coherent with one another as well as with the consequences
of decreased thyroid hormone levels. Because of the coherence across effects on the thyroid and
several interrelated developmental effects in mice (i.e., delays and hormonal changes), the
evidence in animals for developmental effects supports a hazard. There is no reason to expect
that the specific developmental delays observed in mice would not be directly relevant to similar
processes in humans. Thus, based on findings in animals that are presumed to be relevant to
humans, the available evidence supports a hazard and the developing offspring is considered a
potential target for PFBS toxicity in humans. Because no studies in humans were available that
investigated these endpoints, this represents an area deserving of additional research.
5.3 REPRODUCTIVE EFFECTS
Reproductive outcomes, including male and female fertility, pregnancy outcomes,
hormone levels, markers of reproductive development, and reproductive organ weights and
histopathology, have been evaluated in a number of high-confidence studies in mice (Feng et al ..
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2017) and rats (N I P. 2019; Lieder et al.. 2009a; Lieder et al.. 2009b). In addition, five
low-confidence human studies evaluated potential associations between PFBS exposure and
reproductive effects (Yao et al.. 2019; Song et al.. 2018; Zhang et al.. 2018; Zhou et al.. 2017a;
Zhou et al.. 2016).
PFBS exposure has resulted in no significant changes in male mating and fertility
parameters, reproductive organ weights, or reproductive hormones. Although there were some
slight, statistically significant effects on male reproductive endpoints in two rat studies
[specifically, altered sperm parameters such as percentage of abnormal sperm or testicular sperm
count (N I P. 2019; l.ieder et al.. 2009a) and delayed preputial separation at 1,000 mg/kg-day
(l.ieder et al.. 2009a)"I. these findings were observed only at the highest doses and the levels of
change were of questionable biological significance. No significant reproductive effects in men
were noted across two human studies (Song et al.. 2018; Zhou et al.. 2016). although U.S. EPA
noted a nonsignificant inverse association with testosterone and estradiol in male infants in one
study (Yao et al.. 2019).
In general, PFBS exposure in adults has also resulted in no significant alterations in
female fertility or pregnancy outcomes in rats or mice (N I P. 2019; Feng et al.. 2017; l.ieder et
al.. 2009a; l.ieder et al.. 2009b) or in two human studies (Yao et al.. 2019; Zhang et al.. 2018;
Zhou et al.. 2017a; Zhou et al.. 2016). and inconsistent changes in rodent reproductive organ
weights were reported across studies regardless of duration and timing of exposure. However,
changes in normal estrous cyclicity, specifically prolongation of the diestrus stage, have been
reported in both nonpregnant adult rats exposed to PFBS (N I P. 2019) and adult mouse offspring
exposed gestationally from GDs 1 to 20 (Feng et al.. 2017). PFBS exposures in N I P (2019)
began between 8 and 10 weeks of age; although the exposures might overlap with some aspects
of reproductive development or changes in function during adolescence, these rats were sexually
mature and thus the endpoints are considered in the context of reproductive, rather than
developmental, effects. The mouse offspring in the study by Feng et al. (2017) also displayed
delayed vaginal patency and histopathological markers of decreased fertility (i.e., decreased
follicles and corpora lutea); however, the reproductive function of those offspring was not tested.
While adult rat offspring (Fi) in a two-generation toxicity study also exhibited variable changes
in estrous cyclicity (l.ieder et al.. 2009b). including prolonged diestrus at 100 mg/kg-day, this
effect was not observed at higher doses, limiting interpretation, and no effects on vaginal patency
were observed. Female reproductive hormones can inform the potential for effects on
reproductive organ development, estrous cyclicity, and fertility. Changes in serum hormones
included increased testosterone after exposure of female rats as adults ( N I P. 2019). increased
luteinizing hormone and decreased estradiol in pubertal mice after gestational exposure (Feng et
al.. 2017). and decreased estradiol and progesterone when these gestationally exposed mice were
assessed as adults. Overall, the pattern and timing of hormonal changes after PFBS exposure is
difficult to interpret and likely incomplete. However, the hormonal alterations after gestational
PFBS exposure in mice are most relevant to conclusions about female reproductive health.
Taken together, the evidence indicates that the developing reproductive system,
particularly in females, might be a target for PFBS toxicity. However, the potential for
reproductive effects in adults was less clear, and significant impacts on mating or fertility
parameters were not observed across the available studies. Therefore, the evidence in
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developing animals is considered most informative to conclusions relating to potential
developmental effects (see above) and the evidence for reproductive effects (i.e., in adults) is
equivocal. In the three studies of potential reproductive effects in humans, no clear associations
were observed, so the evidence in human studies is equivocal. Overall, based on equivocal
human and animal evidence, the available evidence for reproductive effects is equivocal.
5.4 RENAL EFFECTS
Renal effects associated with oral exposure to PFBS have been observed in adult or
developing rats across high- or medium-confidence gavage studies of various duration ( N I P.
2019; l.ieder et al.. 2009a: l.ieder et al.. 2009b: 3M. 2001. 2000d).
Statistically significant increases in kidney weights have been observed in male and
female rats after short-term exposure in one study (N I P. 2019). with strong dose-dependence for
changes in relative weights in female rats at doses as low as 62.6 mg/kg-day. This study was
likewise the only study to observe changes in serum markers of renal injury, specifically
increased BUN in males at >250 mg/kg-day. However, while several other studies noted slight
increases in weights, typically at higher PFBS doses (>500 mg/kg-day), U.S. EPA found that
these nonsignificant changes were not consistently observed across the set of available studies
and no other studies reported changes in serum markers of renal injury (l.ieder et al.. 2009a:
l.ieder et al.. 2009b: 3M. 2001. 2000d).
Several kidney histopathology lesions (i.e., CPN, hydronephrosis, tubular degeneration,
and tubular dilation) were unaffected by PFBS exposure in rats, although each of these endpoints
was not assessed across several studies (NTP. 2019: l.ieder et al.. 2009a: 3M, 2000d). Mixed
results were reported for mineralization and necrosis. Both of these endpoints were noted in
females, but not males, after subchronic exposure to 600 mg/kg-day (l.ieder et al.. 2009a).
whereas mineralization was unaffected in male or female rats after short-term exposure (3M.
2000d). and necrosis was unaffected in male or female rats in short-term and two-generation (in
both generations) studies (NTP. 2019: l.ieder et al.. 2009b). Multiple markers of inflammatory
changes were consistently noted in the two longest exposure duration studies, which were the
only studies to report on these endpoints. Specifically, increases in chronic pyelonephritis,
tubular basophilia, and mononuclear cell infiltration were observed in female, but not male, rats
following subchronic exposure to 600 mg/kg-day (l.ieder et al.. 2009a). Similarly, increases in
papillary edema and hyperplasia were observed in male and female rats after subchronic
exposure to 600 mg/kg-day (l.ieder et al.. 2009a). and in both generations of rats in the
two-generation study at >300 mg/kg-day (l.ieder et al.. 2009b). with female rats being more
sensitive than males.
Overall, the evidence in animals suggests an increased sensitivity of female rats
(i.e., based on histopathology and organ-weight changes). Due primarily to the consistency and
coherence in renal effects observed in the subchronic study by l.ieder et al. (2009a) and the
reproductive toxicity study by l.ieder et al. (2009b) in male and female rats, the evidence in
animals supports a hazard. There is insufficient evidence in the epidemiology studies of PFBS
to inform the human relevance of these findings. Taken together, the renal histopathology
evidence in rodents identifies a toxicologically significant spectrum of effects that is presumed to
be relevant to similar changes known to occur in humans. Renal effects (i.e., uric acid) were
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evaluated in one low-confidence human study, and no clear association was observed; therefore,
the evidence in human studies is equivocal. Overall, based on findings in animals that are
presumed to be relevant to humans, the available evidence supports a hazard and indicates the
kidney as a target organ of PFBS toxicity.
5.5 HEPATIC EFFECTS
Hepatic effects, including organ-weight changes and histopathology associated with oral
exposures to PFBS, have been observed in high- or medium-confidence studies in adult or
developing rats following short-term- and subchronic-duration exposures (N I P. 2019; Lieder et
al.. 2009a; 3M. 2001. 2000d) and in a two-generation reproductive study in rats (Lieder et al..
2009b). Increased absolute and/or relative liver weights were consistently observed in male and
female rats after short-term and multigenerational exposure (N I P. 2019; l.ieder et al.. 2009b;
3M. 2001. 2000d). In some studies, the magnitude of the liver-weight changes and the doses at
which effects occurred differed across sexes of rat, although the pattern across studies was
unclear and did not consistently indicate one sex as more sensitive. Liver histopathology,
including necrosis and inflammation, was not consistently observed across PFBS studies. One
possible exception is increases in hepatocellular hypertrophy in male rats observed across two
studies (N I P. 2019; l.ieder et al.. 2009b). although female rats were unaffected in the
multigenerational study and this lesion was not observed at up to 600 mg/kg-day in the
subchroni c study by l.ieder et al. (2009a). The only study to observe changes in serum markers
of liver injury was NTP (2019), at >250 mg/kg-day in females and >500 mg/kg-day in males.
The biological relevance or significance of the observed liver effects is not clear. In particular,
the adversity of the variable changes in liver weight and observations of cellular hypertrophy is
unclear. Further, the observed lesions either occurred in only one sex of rat, were not dose
dependent compared with control, and/or occurred only at the highest PFBS dose tested. Thus,
the evidence in animals is equivocal. Overall, based on equivocal animal evidence and a lack of
human studies, the available evidence for hepatic effects is equivocal.
5.6 EFFECTS ON LIPIDS OR LIPOPROTEINS
Few studies have examined the effects of PFBS on circulating or hepatic lipid or
lipoprotein homeostasis. It is recognized that increased circulating levels of lipids and
lipoprotein products and/or increased hepatic lipid load are clinical observations of concern in
humans. However, the lack of effect on lipid dynamics in most studies of rats exposed to high
oral K+PFBS doses for up to 90 days and the generally modest effects in transgenic mice, fed a
high-fat, western-type diet renders this potential health outcome of unclear toxicological
significance at this time. Thus, given the inconsistent, modest effects and the unclear biological
relevance of these changes in isolation (i.e., lipids/lipoproteins were decreased, not increased) the
evidence in animals is equivocal. Effects on serum lipids were evaluated in one low-confidence
human study and childhood adiposity was evaluated in one medium-confidence study. Although
an association was observed between increased PFBS exposure and increased total cholesterol
and higher adiposity, this evidence in humans is equivocal due to lack of additional supportive
evidence. Overall, based on equivocal evidence in both animal and human studies, the available
evidence for effects on lipid or lipoprotein homeostasis is equivocal.
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5.7 IMMUNE EFFECTS
Immune effects were observed in two human studies, including associations with asthma
(Dong et al.. 2013a) and atopic dermatitis (Chen et al.. 2018). Exposure of human peripheral
blood leukocytes or human promyelocytic THP-1 cells to PFBS, in culture, decreased cytokine
(e.g., TNFa and IL-10) secretion following antigen challenge (Corsini et al.. 2012). Because of
the lack of additional evidence and some concerns about potential for residual confounding by
other PFAS, the evidence in human studies is equivocal. Overall, based on equivocal evidence
in human studies and a lack of animal studies, the available evidence for immune effects is
equivocal.
5.8 CARDIOVASCULAR EFFECTS
Cardiovascular effects were observed in two human studies, including associations with
cardiovascular disease in adults (Huang et al.. 2018) and hypertensive disorders in pregnancy
(Huang et al.. 2019b). The results are compelling, but as with the evidence for immune effects,
there is a lack of additional supportive evidence and some concerns about potential for
confounding; thus, the evidence in human studies is equivocal. Overall, based on equivocal
evidence in human studies and a lack of animal studies, the available evidence for cardiovascular
effects is equivocal.
5.9 EVIDENCE INTEGRATION AND HAZARD CHARACTERIZATION
SUMMARY
Based on the evidence integration judgments regarding the potential for PFBS exposure
to cause health effects (the narrative above is summarized in Table 7), the animal studies
informing the potential effects of PFBS exposure on thyroid function, renal function, and
development were concluded to support a hazard. Thus, for the purposes of this assessment, the
animal data supporting these outcomes were considered for use in dose-response analysis, and
other data were considered no further.
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Table 7. Summary of Hazard Characterization and Evidence Integration Judgments
Studies and Confidence
Factors That Increase
Support for Hazard
Factors That Decrease
Support for Hazard
Summary of Findings
Overall
Evidence
Integration
Judgment and
Basis
Thyroid effects
Human studies
Supports a
hazard
(animal evidence
supports a
hazard; human
evidence is
equivocal).
The primary
basis for this
judgment is
thyroid hormone
decreases in mice
and rats at
>62.6 mg/kg-d.
• Low-confidence
case-control study
(Zhang et al.. 2018)
• No factors noted.
• Single study of low
confidence and poor
sensitivity.
No association of PFBS with free T3, free T4, or
thyroid stimulating hormone, but the study had
poor sensitivity and other methodological
limitations that hinder interpretability.
Animal studies (all savage)
Mouse Studies:
•	High-confidence
gestational
(GDs 1-20) exposure
studv (Feng et al..
2017)
Rat Studies:
•	High-confidence
short-term (28-d)
toxicity studv (NTP.
2019)
•	Consistent thyroid
hormone decreases
(i.e., for total T3, total T4,
and free T4) across two
high-confidence studies
of varied design. The
findings were consistent
across two species, sexes,
life stages, and exposure
durations.
•	Dose-response gradients
were observed for those
thyroid hormones.
•	Large magnitudes of
effect (e.g., up to -50%
reductions in offspring
serum hormones) were
reported for those thyroid
hormones.
• No factors noted.
Similar patterns of decreases in thvroid
hormones (i.e.. for total Ti. total T4. and free T4)
were observed in PFBS-exposed pregnant mice
and gestationally exposed female mouse
offspring at >200 mg/kg-d (Feng et al.. 2017)
and in adult female and male rats at
>62.6 mg/kg-d (NTP. 2019).
Increased TSH was reported in mouse dams and
in pubertal (PND 30) offspring following
gestational exposure (Feng et al.. 2017). but no
changes were noted in rats exposed as adults
(NTP. 2019).
Thvroid weight and histonathologv were not
changed after short-term exposure in adult male
or female rats (NTP. 2019).
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Table 7. Summary of Hazard Characterization and Evidence Integration Judgments
Studies and Confidence
Factors That Increase
Support for Hazard
Factors That Decrease
Support for Hazard
Summary of Findings
Overall
Evidence
Integration
Judgment and
Basis
Developmental effects
Human studies
Supports a
hazard
(animal evidence
supports a
hazard; human
evidence is
equivocal).
The primary
basis for this
judgment is a set
of persistent
developmental
delays and
alterations in
reproductive
system
maturation in
female mice,
generally at
>200 mg/kg-d.
No studies available to
evaluate.
--
--
--
Animal studies (all pavape)
Mouse Studies:
• High-confidence
gestational
(GDs 1-20) exposure
studv (Feng et al.
2017s)
•	Biologically consistent
spectrum of
developmental effects in
female offspring in a
high-confidence mouse
study at doses not causing
maternal toxicity,
including pronounced and
persistent effects on B W,
delays in developmental
milestones and sexual
maturation, concordant
effects on reproductive
organs, and altered serum
hormones.
•	Concerning magnitude of
effect (e.g., -25% change
in pup weight) and
dose-dependence for
several parameters.
•	Coherence of effects with
thyroid hormone
insufficiency (see above).
•	Developmental effects
were limited to changes
in one study, sex, and
species.
•	A high-confidence rat
study reported some
inconsistent evidence,
including lack of a
delay in vaginal
patency and lack of
clear effects on estrous
cyclicity or ovarian
morphology, although
the latter endpoint was
assessed in much older
animals. These
potential differences
across species are not
explainable based on
toxicokinetics alone.
In the onlv mouse studv (Feng et al.. 2017).
developmental effects and altered markers of
female reproductive development or function
were observed in female offspring after
gestational PFBS exposure, including decreased
BW. delaved eve ooening. delaved vaginal
Rat Studies:
•	Two high-confidence
gestational exposure
(GDs 6-20) studies:
a range-finding study
and a follow-up
studv (York. 2003c.
2002)
•	High-confidence
two-generation study
(Lieder et al.. 2009b)
onetting, altered estrous cvclicitv (including
prolonged diestrus). altered reproductive
hormones (e.g.. decreased estradiol and
progesterone), and effects on reproductive
organs (e.g.. weight and ovarian morphology).
Most effects were observed at >200 mg/kg-d,
with several changes noted at PND 60.
Endpoints relating to fertility, oregnancv.
survival, and fetal alterations were unchanged
in both rats and mice across the four available
studies, although this was not tested in mouse
offspring (Feng et al.. 2017).
Developmental BW changes in rat offspring
were either unchanged (Lieder et al.. 2009b) or
observed only at doses causing parental toxicity
(York. 2003c. 2002).
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Table 7. Summary of Hazard Characterization and Evidence Integration Judgments
Studies and Confidence
Factors That Increase
Support for Hazard
Factors That Decrease
Support for Hazard
Summary of Findings
Overall
Evidence
Integration
Judgment and
Basis
Continued:
Continued:
Note: these effects were also
coherent with effects on estrous
cyclicity observed after
short-term exposure in adult rats
(NTP. 20191 but this was
categorized as a reproductive
effect (see below).
Continued:
Continued:
In a rat two-generation study, while some
statistically significant findings were noted for
markers of female reproductive development or
function, thev were not dose-dependent or were
of questionable biological relevance; thus, no
clear changes in Fi offspring were noted at
doses up to 1.000 mg/kg-d regarding vaginal
patencv or estrous cvcling at comparable ages
to (Feng et al.. 2017). or in ovarian morphologv
after the Fi females gave birth to the F2 pups.
Continued:
Reproductive effects
Human studies

Male reproductive effects
•	Low-confidence
cohort studv (Zhou et
al. 2016s)
•	Low-confidence
cross-sectional study
(Song et al.. 2018)
•	Low-confidence
cross-sectional study
(Yao et al.. 2019)
• No factors noted.
• Lack of clear
association in studies of
low confidence with
poor sensitivity
(i.e., due to low
exposure levels, range).
No clear association between PFBS exposure
and male reproductive hormones (Zhou et al..
2016) or semen parameters (Song et al.. 2018)
in adults. A study in newborns reported
nonsignificant inverse associations between
PFBS exposure and testosterone and estradiol
(Yao et al.. 2019).
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Table 7. Summary of Hazard Characterization and Evidence Integration Judgments
Studies and Confidence
Factors That Increase
Support for Hazard
Factors That Decrease
Support for Hazard
Summary of Findings
Overall
Evidence
Integration
Judgment and
Basis
Female reproductive effects

•	Low-confidence
cross-sectional study
(Zhou et al. 20173)
•	Low-confidence
cohort studv (Zhou et
al.. 2016s)
•	Low-confidence
cross-sectional study
(Yao et al.. 2019)
•	Low-confidence
case-control study
(Zhang et al.. 2018)
• No factors noted.
•	Lack of clear
association in studies of
low confidence with
poor sensitivity
(i.e., due to low
exposure levels, range).
•	Potential for reverse
causation for menstrual
cycle characteristics
and premature ovarian
insufficiency.
No clear association between PFBS exposure
and female reproductive hormones (Zhou et al..
2016) or menstrual cvcle characteristics (Song
etal.. 2018).
Equivocal
(equivocal human
and animal
evidence).
Note: As the
strongest
evidence for
female
reproductive
Animal studies (all savage)
effects was in
offspring that
were
Male reproductive effects
Rat Studies:
•	High-confidence
short-term (28-d)
toxicity studv (NTP.
2019)
•	High-confidence
two-generation study
(l.ieder et al.. 2009b)
•	High-confidence
subchronic study
(l.ieder et al.. 2009a)
• No factors noted.
•	A few small,
statistically significant
changes were not
dose-dependent or were
of questionable
biological relevance.
•	Lack of effects on male
mating and fertility,
hormones, or
reproductive organs in
rats.
Statistically significant effects on sperm health
(NTP. 2019; Lieder et al.. 2009a) and delayed
preputial separation at 1.000 mg/kg-d (Lieder et
gestationally
exposed, these
findings were

al.. 2009b) were not observ ed at lower doses,
were within the normal range of historical
controls for the laboratory, and/or were no
longer significantly changed after correcting for
other variables (e.g., BW).
Other relev ant parameters (e.g.. organ weights,
mating success, and so forth) were unchanged
in the three studies.
considered most
relevant to
developmental,
not reproductive,
effects.
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Table 7. Summary of Hazard Characterization and Evidence Integration Judgments
Studies and Confidence
Factors That Increase
Support for Hazard
Factors That Decrease
Support for Hazard
Summary of Findings
Overall
Evidence
Integration
Judgment and
Basis
Female reproductive effects

Mouse Studies:
•	High-confidence
gestational
(GDs 1-20) exposure
studv (Feng et al.
2017)
Rat Studies:
•	High-confidence
short-term (28-d)
toxicity studv (NTP.
2019)
•	High-confidence
subchronic study
(l.ieder et al.. 2009a)
•	High-confidence
two-generation study
(l.ieder et al.. 2009b)
•	Effects on markers of
female reproductive
function (i.e., estrous
cyclicity) were observed
in high-confidence
studies in rats and mice.
•	Changes in reproductive
serum hormones were
observed in female rats
(i.e., increased
testosterone) and mice
(e.g., decreased estradiol
and progesterone).
Although the pattern of
change is difficult to
interpret and likely
incomplete, there were no
conflicting data.
•	Lack of similar effects
on reproductive
function (i.e., estrous
cyclicity) in a second
high-confidence rat
study.
•	Lack of effects on
female fertility or
pregnancy measures,
although this was
untested in prenatally
exposed female mouse
offspring.
•	Lack of organ-weight
changes in three rat
studies.
Note: The lack of effects on
ovarian follicles in rats did
not decrease the support for
hazard provided by findings
in mice, as the age at endpoint
assessment was not
comparable.
See "Developmental effects" (above) for
findings from Feng et al. (2017) and Lieder et
al. (2009b).
Altered estrous cvclicitv (including prolonged
diestrus) and increased serum testosterone were
observed in female rats after short-term
exposure, primarily at >250 mg/kg-d (NTP.
2019).
Female reproductive organ weights were
reduced in gestationally exposed mouse
offspring (Feng et al.. 2017). but were
unchanged after short-term, subchronic, or
two-generational exposure (NTP. 2019; Lieder
et al.. 2009a: Lieder et al.. 2009b).

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Table 7. Summary of Hazard Characterization and Evidence Integration Judgments
Studies and Confidence
Factors That Increase
Support for Hazard
Factors That Decrease
Support for Hazard
Summary of Findings
Overall
Evidence
Integration
Judgment and
Basis
Renal effects
Human studies
Supports a
hazard.
(animal evidence
supports a
hazard; human
evidence is
equivocal).
The primary
basis for this
judgment is
kidney
histopathology in
rats, primarily
females, at
>300 mg/kg-d.
• Low-confidence
cross-sectional study
(OinetaL 2016s)
• No factors noted.
•	Inconsistency across
subpopulations in
single study.
•	Single study of low
confidence with
concern for potential
reverse causality.
Overall, there was no clear association for
PFBS and uric acid. No association observed
between PFBS and uric acid in the total
population. Increase in uric acid with increased
exposure in boys but decrease for girls (neither
was statistically significant).
Animal studies (all savage)
Rat Studies:
•	One high-confidence
subchronic study
(l.ieder et al. 20093)
•	Two high-confidence
studv (NTP. 2019;
3M. 2001s) and one
medium-confidence
(3M. 2000d)
short-term (10-28 d)
study
•	One high-confidence
two-generation study
(l.ieder et al.. 2009b)
•	Two high-confidence
studies with the longest
exposure durations
reported consistent effects
on kidney histopathology
in male and female rats
(females were more
sensitive).
•	The histopathological
effects related to
inflammation were
largely dose-dependent
and of a concerning
magnitude, although
primarily at high doses
(300 or 600 mg/kg-d).
•	Inconsistency in
kidney-weight changes
across studies.
•	Findings are from a
single laboratory and
species.
Note: The general lack of
effects on other pathology
endpoints in the shorter term
studies was not considered to
decrease support for hazard,
as this was not interpreted as
inconsistent.
Increases in kidttev weight in male and female
rats were observed in one short-term study at
>62.6 mg/kg-d, but clear changes were not
observed in the other short-term, subchronic, or
two-generation rat studies.
Kidttev histonatholoev for some effects
(i.e., CPN, hydronephrosis, tubular
degeneration, and tubular dilation) was
unchanged in single-study evaluations, and
mixed results across studies were reported for
mineralization and necrosis (NIP. 2019; l.ieder
et al.. 2009a: Lieder et al.. 2009b: 3M. 2000d).
Multiple markers potentially related to
inflammation and most notably capillary edema
and hyperplasia were increased in the two
longest duration studies (Lieder et al.. 2009a:
Lieder et al.. 2009b). without contrary evidence.
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EPA/690/R-21/001F
Table 7. Summary of Hazard Characterization and Evidence Integration Judgments
Studies and Confidence
Factors That Increase
Support for Hazard
Factors That Decrease
Support for Hazard
Summary of Findings
Overall
Evidence
Integration
Judgment and
Basis
Continued:
Continued:
Continued:
Continued:
Other markers of renal iniurv. including BUN
and creatinine, were mostly unaffected across
studies (NTP. 2019: Lieder et al.. 2009a: Lieder
etal.. 2009b: 3M. 2001. 2000d). although the
NTP study did observe effects on BUN in males
at >250 mg/kg-d.
Continued:
Hepatic effects
Human studies

No studies available to
evaluate
-
-
-

Animal studies (all savage)

Rat Studies:
•	One high-confidence
subchronic study
(l.ieder et al. 20093)
•	Two high-confidence
studies (NTP. 2019:
3M. 2001s) and one
medium-confidence
(3M. 2000d)
short-term (10-28 d)
study
•	One high-confidence
two-generation study
(l.ieder et al.. 2009b)
• Consistent changes in
liver weights in rats of
both sexes across four
studies. Although the
pattern (e.g., by sex and
dose) and magnitude of
changes varied across
studies, weights were
consistently increased.
•	Other than liver-weight
changes, there were
notable unexplained
inconsistencies in the
findings across studies.
•	One high-confidence
study was entirely
inconsistent.3
Absolute or relative liver weights were
increased in all studies except the 90-d exposure
component of the studv bv Lieder et al. (2009a).
which tested doses up to 600 mg/kg-d.
Note: 70 d of exposure in this study did elicit
effects.
Effects generally occurred at >300 mg/kg-d,
although one study reported effects at lower
doses (NTP. 2019: 3M. 2001). and two others
(3M. 2001. 2000d) observed changes at
>900 mg/kg-d.
Serum markers of liver iniurv were unchanged
Equivocal
(equivocal human
and animal
evidence).


in three studies (Lieder et al.. 2009a: 3M. 2001.
2000d) and increased in one short-term studv at
>250 mg/kg-d (NTP. 2019).

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Table 7. Summary of Hazard Characterization and Evidence Integration Judgments
Studies and Confidence
Factors That Increase
Support for Hazard
Factors That Decrease
Support for Hazard
Summary of Findings
Overall
Evidence
Integration
Judgment and
Basis
Continued:
Continued:
Continued:
Continued:
Liver histonatholoev. specifically
hepatocellular hypertrophy and cytoplasmic
alterations in males and females (NTP. 2019) or
hvDertroDhv in females onlv (Lieder et al..
2009a). were noted in two studies, but not in the
others.
Continued:
Lipid or lipoprotein homeostasis
Human studies
Equivocal
{equivocal human
and animal
evidence).
•	Low-confidence
cross-sectional study
(Zernr et al.. 2015)
•	Medium-confidence
studv (Chen et al..
2019)
•	Statistically significant
association in
medium-confidence study
of adiposity.
•	Exposure response
gradient observed across
tertiles for adiposity.
•	Single study per
outcome.
•	Potential for residual
confounding.
Increase in total cholesterol (statistically
significant, (3: 19.3 mg/dL increase per unit
increase in PFBS) (Zernr et al.. 2015). Hisher
adiposity in 5-year-old children associated with
hisher levels of PFBS in cord blood (Chen et
al.. 2019).
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Table 7. Summary of Hazard Characterization and Evidence Integration Judgments
Studies and Confidence
Factors That Increase
Support for Hazard
Factors That Decrease
Support for Hazard
Summary of Findings
Overall
Evidence
Integration
Judgment and
Basis
Animal studies

Mouse Studies (diet):
• Medium-confidence
short-term (4-6 wk)
studv (Biilattd et al..
2011); transgenic
mice (human-like
lipid metabolism)
were fed a high-fat
diet
Rat Studies (all eavaee):
• Decreases in serum
cholesterol and
triglycerides were
observed in male rats and
mice.
•	Inconsistent evidence
in other rat studies and
across sexes.
•	Small effect
magnitudes and unclear
direction (decreases) of
changes are of
questionable biological
relevance and could not
be informed by
evaluating
dose-dependency
(i.e., only single-dose
or high-dose effects
were observed).
Serum lipids, specifically cholesterol and
triglyceride levels, were slightly decreased
(-20%) at 900 mg/kg-d in males, but not
females, in one rat studv (3M. 2001). but not in
two other rat studies at up to 1,000 mg/kg-d.
Serum and hepatic lipids and lipoproteins were
also decreased in male mice exposed to
~30 mg/kg-d in diet.
•	One high-confidence
subchronic study
(Lieder et al.. 2009a)
•	One high-confidence
studv (3M. 2001) and
one
medium-confidence
(3M. 2000d)
short-term (10-28 d)
study
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Table 7. Summary of Hazard Characterization and Evidence Integration Judgments
Studies and Confidence
Factors That Increase
Support for Hazard
Factors That Decrease
Support for Hazard
Summary of Findings
Overall
Evidence
Integration
Judgment and
Basis
Immune effects
Human studies

Asthma

• Medium-confidence
case-control study
(Zhou et al. 2016;
Zhu et al.. 2016;
Done et al.. 2013b)
• Statistically significant
association in a
medium-confidence
study.
Note: Increases in eosinophil
markers were not interpreted to
increase support for hazard,
because they were not
statistically significant and other
markers important to asthma
etiology (e.g., IgE) were
unchanged.
• Association was
observed in a single
study with concern
regarding the potential
for residual
confounding (e.g., with
other PFAS chemicals).
Statistically significant increase in odds of
asthma diagnosis in the previous year
(OR: 1.2-1.9) with increased PFBS exposure.
Eosinophil markers (i.e., AEC and ECP) were
increased with increased PFBS exposure in
asthmatics and nonasthmatics; however, these
increases did not reach statistical significance.
IgE and T-helper cell-specific cytokines were
unchanged (Zhu et al.. 2016).
Equivocal
{equivocal human
and animal
evidence).
Atopic dermatitis

• Medium-confidence
cohort studv (Chen et
al.. 2018s)
• No factors noted.
• Slight associations
were not statistically
significant in a single
study with concern
regarding the potential
for residual
confounding (e.g., with
other PF AS chemicals).
Statistically nonsignificant increase in odds of
atopic dermatitis (OR: 1.2) with increased
PFBS exposure.

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Table 7. Summary of Hazard Characterization and Evidence Integration Judgments
Studies and Confidence
Factors That Increase
Support for Hazard
Factors That Decrease
Support for Hazard
Summary of Findings
Overall
Evidence
Integration
Judgment and
Basis
Animal studies

No studies available to
evaluate.
-
-
-
Cardiovascular effects
Human studies
Equivocal
{equivocal human
and animal
evidence).
•	Medium-confidence
cross-section study
(Huang et al. 2018)
•	Medium-confidence
cross-sectional study
(Huang et al.. 201%)
• Statistically significant
associations in
medium-confidence
studies.
• Single study per
outcome.
Higher odds of cardiovascular disease (total and
individual types of disease) with PFBS
exposure (Huang et al.. 2018). Higher odds of
hypertensive disorders in pregnancy with higher
PFBS exposure (Huang et al.. 2019b). There is
potential for residual confounding that
decreases confidence in the evidence.
Animal studies
No studies available to
evaluate.
-
-
-
aThe lack of liver effects in the subchronic study was not interpreted to significantly reduce support for hazard because the maximum tolerated dose was
600 mg/kg-d, and other studies reported only liver effects at >900 mg/kg-d.
AEC = absolute eosinophil count; BUN = blood urea nitrogen; BW = body weight; CPN = chronic progressive nephropathy; ECP = eosinophilic cationic protein;
GD = gestation day; IgE = immunoglobulin E; NTP = National Toxicology Program; OR = odds ratio; PFAS = per- and polyfluoroalkyl substances;
PFBS = perfluorobutane sulfonic acid; PND = postnatal day; T3 = triiodothyronine; T4 = thyroxine; TSH = thyroid stimulating hormone.
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6.0 DERIVATION OF VALUES
The hazard and dose-response database for PFBS and the potassium salt is primarily
associated with the oral route of exposure. There are a limited number of dermal studies (see
Table 5) and no known inhalation studies. There are no known studies evaluating potential
cancer effects of PFBS. Therefore, only noncancer reference values are derived in this
assessment for the oral route.
6.1 DERIVATION OF ORAL REFERENCE DOSES
The hazards of potential concern for oral PFBS exposure include thyroid, developmental,
and kidney effects. Overall, the evidence supports a hazard for thyroid, developmental, and
kidney effects based on the evidence from animal studies. The limited evidence for thyroid or
renal effects in human studies is equivocal, and no studies evaluating developmental effects
following PFBS exposure in humans were available. Thus, data in humans were not considered
further, and the available animal studies that evaluated these effects are considered in the
derivation of oral RfDs.
6.1.1 Derivation of the Subchronic Oral Reference Dose
6.1.1.1 Estimation of Points of Departure
Effects in the thyroid were considered when determining potential PODs for deriving a
subchronic RfD. Similar patterns of decreases in total T.<. total T-t. and free T-t were observed in
PFBS-exposed pregnant mice, nonpregnant adult female rats, adult male rats, and gestationally
exposed female mouse offspring (N I P. 2019; Feng et al.. 2017). These decreases were
significant (-20% in dams and -50% in offspring), were shown to persist at least 60 days after
gestational exposure in offspring, and they exhibited a clear dose dependence in both studies.
Reflex increases in TSH in response to decreased T4 or T3 were not observed in male or female
rats following 28 days of exposure (NTP. 2019). Such an increase in TSH was observed in
pregnant mice (measured at GD 20) and their corresponding female offspring, at PND 30 only,
with an irregular dose-response or time course (Feng et al .. 2017). This pattern of decreased
thyroid hormone without a concomitant increase in TSH is consistent with a human clinical
condition referred to as "hypothyroxinemia" (Negro et al.. 2011). Importantly, milder forms of
thyroid perturbation are up to 10 times more prevalent in human populations than overt
gestational hypothyroidism (Korcvaar et al.. 2016; Stagnaro-Green et al.. 201 1).
Hypothyroxinemia has been associated with impairments in neurodevelopment and/or cognition
later in life (Thompson et al.. 2018; Min et al.. 2016). Because the single available study in
humans had severe limitations hindering the interpretation of the relationship between PFBS
exposure and thyroid hormone alterations, at this time the available evidence in humans is not
able to inform the potential for thyroid effects in humans. This hypothyroxinemia, rather than
overt or subclinical hypothyroidism, is further supported by the lack of effect on thyroid weight
or tissue architecture in rats after 28 days of PFBS exposure (N I P. 2019).
Developmental effects were considered in determining potential PODs for derivation of a
subchronic RfD. Specifically, in Feng et al. (2017). developmental delays or abnormalities in
growth (i.e., BW and eye opening), reproductive organs (i.e., ovaries, uterus, and vaginal
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opening), and reproductive cycling (i.e., first estrous and prolongation of diestrus) were observed
in mouse offspring. These effects were observed in mice from litters in which thyroid hormone
deficiency occurred at PND 1 and then sustained through pubertal and adult periods
(i.e., PNDs 30 and 60, respectively). These interrelated developmental effects in mice
(i.e., delays and hormonal changes) are coherent with effects on the thyroid and presumed to be
directly relevant to similar processes in humans; however, studies evaluating these outcomes in
humans are not available.
Effects in the kidney were considered in determining potential PODs for deriving a
subchronic RfD. Lieder et al. (2009a) reported mild to moderate hyperplasia in the kidneys of
male and female rats following subchronic-duration exposure to PFBS, and l.ieder et al. (2009b)
found the same effects in the Po- and Fi-generation animals in their reproductive toxicity study.
Other studies evaluating effects in the kidney were of shorter duration and thus less suitable as a
candidate principal study. Additional histopathological alterations accompanied the hyperplasia
observed in the kidney, including papillary edema and inflammatory changes, specifically
increases in chronic pyelonephritis, tubular basophilia, and mononuclear cell infiltration (l.ieder
et al.. 2009a; l.ieder et al.. 2009b). Across the reported kidney histopathological effects
following PFBS exposure, female rats were generally more sensitive than males.
Selected data sets from studies with multiple exposure levels for thyroid, developmental,
and kidney effects were modeled using the U.S. EPA's Benchmark Dose Software (BMDS)
Version 2.7. Consistent with the U.S. EPA's Benchmark Dose Technical Guidance Document
(U.S. EPA. 2012). the BMD and 95% lower confidence limit on the BMD (BMDL) were
estimated using a benchmark response (BMR) to represent a minimal, biologically significant
level of change. Based on BMD guidance, in the absence of information regarding the level of
change that is considered biologically significant, a BMR of 1 SD from the control mean for
continuous data or a BMR of 10% extra risk for dichotomous data is used to estimate the BMD
and BMDL, and to facilitate a consistent basis of comparison across endpoints, studies, and
assessments. For some types of effects (e.g., frank effects, developmental effects), biological
considerations may warrant the use of a BMR of 0.5 SD or lower.
For effects in developing offspring, including thyroid hormone changes, a BMR of
0.5 SD change from the control mean is used for continuous data to account for effects occurring
in a sensitive life stage. A 1 SD BMR is also presented as the basis for model comparison as
directed in the U.S. EPA Benchmark Dose Technical Guidance (U.S. EPA. 2012).
For thyroid hormone effects in offspring, a biological level of concern was considered in
the identification of a BMR. Multiple lines of evidence regarding the degree of thyroid hormone
disruption and developmental outcomes in offspring were evaluated. During developmental life
stages, such as gestational/fetal and postnatal/early newborn, thyroid hormones are critical in
many physiological processes associated with somatic growth and maturation and with life
functions like thermogenesis, pulmonary gas exchange, and cardiac development (Sferruzzi-Perri et
al.. 2013; Hittman et al.. 2012). Further, thyroid hormones are critically important in early
neurodevelopment because they directly influence neurogenesis, synaptogenesis, and myelination
(Rovet. 2014; Puig-Domingo and Vila. 2013; Stenzel and Huttner. 2013; Patel et al.. 2011). Note
that evidence from human epidemiological studies examining the association between thyroid
hormone economy in pregnant mothers and neurodevelopment in their offspring is inconsistent.
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Several human epidemiologic studies have demonstrated key relationships between decreased
levels of thyroid hormones such as FT4 in a pregnant woman and in utero and early postnatal life
neurodevelopmental status. For example, children born euthyroid but who were exposed to
thyroid hormone insufficiency in utero (e.g., <10th percentile free T4), present with cognitive
impairments (e.g., decreased intelligence quotient [IQ], increased risk of expressive language)
and/or concomitant abnormalities in brain imaging (Levie et al.. 2018; Korevaar et al.. 2016;
Henrichs et al.. 2010; Lavado-Autric et al.. 2003; Mirabel la et al. 2000). Maternal
hypothyroxinemia was also associated with adverse motor function and teacher-reported problems
of behavior in offspring at 5 years of age (Andersen et al.. 2018). Other human epidemiologic
studies have not reported significant associations between thyroid hormone status during pregnancy
and neurodevelopmental outcomes in offspring. For example, there was no statistically significant
association between thyroid status and IQ decrements or neuropsychological parameters in children
born to mothers screened and diagnosed with subclinical hypothyroidism (Hales et al.. 2018;
Lazarus et al.. 2012) or mothers undergoing treatment for hypothyroxinemia during gestation
(Casey et al .. 2017). In these studies, the timing of maternal hypothyroxinemia during pregnancy
may be a critical consideration for developmental health outcomes in offspring. Studies have
observed a relationship between low free T4 levels in women at 12 weeks gestation, but not
32 weeks gestation, and impaired psychomotor development in their offspring (Kooistra et al..
2006; Pop et al.. 2003). In addition, differences in the type of maternal disruption of thyroid
homeostasis may affect the interpretation of the human epidemiologic study results. Specifically,
aside from overt primary hypothyroidism, there are two primary subcategories of hypothyroidism:
(1) subclinical hypothyroidism; and (2) hypothyroxinemia. Subclinical hypothyroidism is
characterized by elevated TSH levels with normal serum T4 and T3 concentrations. In contrast,
hypothyroxinemia is characterized by decreased T4 with normal serum concentrations of TSH and
T3 (Alexander et al.. 2017; Choksi et al.. 2003). Maternal T4 is the primary source of thyroid
hormone for a developing human fetus in the first trimester (i.e., little if any maternal T3 is
transferred across the placenta primarily due to high levels of deiodinase 3 activity that catabolizes
T3 to a biologically inactive form). The first trimester is also a critical window for central nervous
system development (e.g., neural tube, spinal cord, medulla, pons, thalamus/hypothalamus, etc.). It
therefore stands to reason that the health implications may be different for early in utero
development if associated with a condition where maternal T4 (and T3) concentrations are normal
(subclinical hypothyroidism) versus one involving decreased levels of T4 (hypothyroxinemia).
With regard to what level of decrease in thyroid hormone (e.g., T4) is sufficient for
anatomical and/or functional alterations, particularly in neurodevelopment in fetuses or
newborns, several studies have identified a range of T4 decrements associated with
neurodevelopmental health outcomes across humans or experimental rodents. For example,
neurodevelopmental and cognitive deficits have been observed in children who experienced a
25% decrease in maternal T4 during the second trimester in utero (Haddow et al.. 1999). In other
studies, mild to moderate thyroid insufficiency in pregnant women was defined as having serum
T4 levels below the 10th percentile for the study population, which was associated with a 15-30%
decrease relative to the corresponding median (Finken et al.. 2013; Julvez et al.. 2013; Roman et
al.. 2013; Henrichs et al.. 2010). In experimental animals, decreases in mean maternal T4 levels
of-10-17%) during pregnancy and lactation have been found to elicit neurodevelopmental
toxicity in rat offspring (Gilbert et al.. 2016; Gilbert. 2011). With regard to a general diagnostic
criterion to delineate hypothyroxinemia from other types of clinical hypothyroidism, the
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Controlled Antenatal Thyroid Study (CATS), conducted in a large cohort of pregnant women in
Europe, resulted in the identification of a condition referred to as "isolated hypothyroxinemia"
and is defined as the presence of free thyroxine (FT4) below the 2.5th percentile with a
thyrotropin (TSH) level within the reference range (Hales et al.. 2018; Lazarus et al.. 2012;
Negro et al.. 2011). However, there is no clear or consistent biological threshold for T4 changes
specifically associated with untoward developmental health outcomes, so a BMR of 0.5 SD was
therefore identified as a default when performing BMD modeling on thyroid hormone alterations
in offspring, consistent with U.S. EPA Benchmark Dose Technical Guidance (U.S. EPA. 2012).
Further, while total T4 (TT4), free T4 (FT4), and TSH dose-response data are BMD modeled (see
Table 9), important biological considerations are presented in Section 6.1.1.2 that delineate TT4
as the key hormone metric for a developing fetus/neonate.
Significantly decreased thyroid hormone (e.g., T4 and T3) was observed in adult rats
exposed twice daily to oral K PFBS (NTP. 2019) for 28-days, as well as the Po (maternal) mice
of the Feng et al. (2017) study. No overt signs of traditional hypothyroidism such as increased
TSH and increased thyroid tissue weight or histopathology were observed in either adult
population. Adult rodents have a considerable reserve thyroid hormone capacity compared with
the developing offspring, which depend on their supply from maternal T4. While there is
concern over decreases in thyroid hormone (i.e., hypothyroxinemia) in developmental life stages
due to critical endocrine dependency of in utero and neonatal development, the levels at which
there is concern for hypothyroxinemia in euthyroid adults is unclear. Therefore, for euthyroid
adult rats and mice, a biologically significant level of change was not determined for the BMR
because it is unclear what magnitude of hormone perturbation would be considered adverse.
Therefore, for thyroid hormone effects in adult rodents, a default BMR of 1 SD from control
mean was applied. Section 6.1.1.2 presents critical distinctions between perturbations in thyroid
hormone economy in adults versus developing fetus/neonates, resulting in the use of different
BMRs across life stages (e.g., 1 SD for adults, 0.5 SD for newborns).
For kidney hyperplasia data from the sub chronic study by l.ieder et al. (2009a) and the
two-generation reproductive toxicity study by l.ieder et al. (2009b). a BMR of 10% extra risk
was used because it is the recommended approach for dichotomous data in the absence of
information on the minimally significant level of change.
6.1.1.2 Approach for Animal-Human Extrapolation of Perfluorobutane Sulfonic Acid
Dosimetry
As discussed in Section 1.3, toxicokinetic data exists for PFBS in relevant animal species
(i.e., rats and mice) and humans, such that a data-informed adjustment approach for estimating
the dosimetric adjustment factor (DAF) can be used. In Recommended Use of Body Weight3/4 as
the Default Method in Derivation of the Oral Reference Dose (U.S. EPA. 201 lb), the U.S. EPA
endorses a hierarchy of approaches to derive human equivalent oral exposures using data from
laboratory animal species, with the preferred approach being physiologically based toxicokinetic
modeling. Other approaches might include using chemical-specific information, without a
complete physiologically based toxicokinetic model. In the absence of chemical-specific models or
data to inform the derivation of human equivalent oral exposures, the U.S. EPA endorses BW3'4
as a default to extrapolate toxicologically equivalent doses of orally administered agents from all
laboratory animals to humans for the purpose of deriving an RfD under certain exposure conditions.
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EPA 690 R-21 00IF
The U.S. EPA concluded that data for PFBS are adequate to support derivation of
data-informed dosimetric adjustment. Briefly, the ratio of the clearance (CL) in humans to
animals, CLu.CLa, can be used to convert an oral dose-rate in experimental animals (mg/kg-day)
to a human equivalent dose rate. Assuming the exposure being evaluated is low enough to be in
the linear (or first order) range of clearance, the average blood concentration (Cavg) that results
from a given dose is calculated as:
r fm§/ ^ - /abs X dose (mg/kg/hr) /
'mU	/ CL (mL/kg/hr)
where/abs is the fraction absorbed and dose is the average dose rate expressed at an
hourly rate. Assuming equal toxicity given equal Cavg in humans as in mice or rats, and that /abs
is the same in humans as animals, the equitoxic dose, human equivalent dose (HED) (i.e., the
human dose that should yield the same blood concentration (Cavg) as the animal dose from
which it is being extrapolated), is then calculated as follows:
POD	CLh
HED = 77;	= POD x —-
cla/	CLa
/CLh
Thus, the DAF could be calculated as simply CLh.CLa, the ratio of clearance in humans
to clearance in the animal from which the POD is obtained. However, clearance values are not
reported for humans in the available toxicokinetic studies for PFBS (Xu et al., 2020; 01 sen et al.,
2009). Because clearance is a measure of average elimination, to calculate clearance in the
absence of the information, one also needs to evaluate a companion variable, the Vd. Neither
Olsen et al. (2009) nor Xu et al. (2020) reported the Vd for humans. However, there is evidence
suggesting that Vd for PFBS is relatively similar across species, including rodents
(e.g., 0.12-0.29 L/kg across male and female rats following 10 mg/kg i.v. dose) and monkeys
(e.g., 0.21-0.25 L/kg across male and female cynomolgus macaques following 10 mg/kg i.v.
dose) (Chengelis et al.. 2009: Olsen et al.. 2009). Therefore, it is reasonable to assume Vd for
humans is approximately equivalent to Vd for animals (i.e., Vd.H = Fd. \), in which case clearance
and half-life are inversely related as follows:
CL (mL/kg/hr) = ln(2) x	x Fd (mL/kg)
^1/2'
Because reliable measures of half-life in humans and animals are available for PFBS, the
ratio of elimination half-life in animals from which the POD is obtained to that in humans,
h/2,A:h/2H, can be used to calculate the DAF, and the human equivalent dose (HED) can be
calculated as follows:
tl/2
HED = POD X	^
fl/2 H
As described in Section 1.3, two studies evaluated the elimination of human serum
K+PFBS in human populations with previous occupational exposure (Xu et al.. 2020: Olsen et
al., 2009). Initial blood concentrations of PFBS in the population examined by Xu et al. (2020)
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EPA/690/R-21/001F
are more representative of environmental exposure, and the population was larger, including 11
male and 6 female employees when compared to 01 sen et al. (2009). While the estimated serum
half-life of PFBS reported by 01 sen et al. (2009) overlapped with that by Xu et al. (2020)
(mean: 43.8 days; range: 21.9-87.6 days), there is a statistically significant difference between
these two studies. As such, the two data sets will not be combined and the half-life estimated by
Xu et al. (2020) is presumed to better predict human dosimetry at environmental levels. The
average half-life reported by Xu et al. (2020) (mean: 43.8 days or 1,050 hours) was assigned for
t\iH.
One study evaluated the elimination of serum PFBS in mice. Lau et al. (2020) reported
serum terminal half-lives of 5.8 hours in male mice and 4.5 hours in female mice. Because the
half-life estimates did not vary significantly between the doses (i.e., 30 and 300 mg/kg), these
parameter estimates were combined. However, there was a statistically significant difference in
the half-life estimates between sexes (female mice had a slightly shorter half-life [4.5 hours]
compared to males [5.8 hours]), so sex-specific half-lives were assigned for fr/2,Afor mice.
Two studies were used to calculate serum half-life estimates for dosimetric adjustment in
rats (Huang et al.. 2019a: 01 sen et al.. 2009). A numerical average of the terminal half-lives
(tv2,p) measured in rats after oral and i.v. doses is identified in 01 sen et al. (2009) as 4.6 hours in
males and 5.7 hours in females. 01 sen et al. (2009) reported sex-specific elimination differences
in half-life values in rats. A numerical average of the tv2,p measured in male rats after oral and
i.v. doses in Huang et al. (2019a) is 4.9 hours. In male rats, half-life values reported in 01 sen et
al. (2009) and Huang et al. (2019a) are consistent, thus they were averaged for use in dosimetric
adjustment, resulting in a geometric mean terminal serum half-life of 4.8 hours. The terminal
half-life value reported by Huang et al. (2019a) in female rats after a 4-mg/kg i.v. dose of PFBS
was 0.95 hours. Huang et al. (2019a) was not able to fit the data to a two-compartment model,
thus they did not report a tv2,p for rats following oral exposure. For this reason, the mean female
tvifi value from 01 sen et al. (2009) was used for dosimetric adjustment.
Table 8 presents the DAFs for converting rat and mice PODs to HEDs for PFBS.
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Table 8. Mouse, Rat, and Human Half-Lives and Data-Informed DAFs
Species
Sex
Animal tin (hr)
Human tvi (hr)
DAF (?i/2,\/?i/2,n)
Mouse
Male
5.8a
l,050b
0.0055
Female
4.5°
0.0043
Rat
Male
4.8d
0.0046
Female
5.T
0.0054
"Terminal serum half-life of combined doses for male mice from Lau et at (2020).
' Mean serum elimination half-life for humans (combined sexes) from Xu et al. (2020).
"Terminal serum half-life of combined doses for female mice from Lau et al. (2020).
'Geometric mean of terminal serum half-lives (?i/2,p) measured after all oral and i.v. doses for male rats from Olsen
et al. (2009) and Huang et al. (2019a).
"Mean of terminal serum half-lives (h/2 p) measured after oral and i.v. doses for female rats from Olsen et al.
(2009).
DAF = dosimetric adjustment factor; i.v. = intravenous; ti/2 = half-life.
Where modeling was feasible, the estimated BMDLs were identified as PODs
(summarized in Table 9). Further details, including the modeling output and graphical results for
the model selected for each endpoint, can be found in HAWC and are discussed in Appendix F.
Where dose-response modeling was not feasible, NOAELs or LOAELs were identified
(summarized in Table 9).
Table 9. PODs Considered for Deriving the Subchronic RfD for
K+PFBS (CASRN 29420-49-3)
Endpoint/Reference
Species/Life
Stage—Sex
POD (HED)a
(mg/kg-d)
Comments*
Thyroid effects
Total Ti—Feng et al. (2017)
Mouse/Po—female
BMDL1SD = 0.093
Adequate model fit
Free T4—Feng et al. (2017)
Mouse/Po—female
NOAEL = 0.21
No models provided adequate
statistical or visual fit to mean
responses
TSH—Feng etal. (2017)
Mouse/Po—female
NOAEL = 0.21
No models provided adequate
statistical or visual fit to mean
responses
Total T4 PND 1 (fetal ri)h—Feng et
al. (2017)
Mouse/Fi—female
NOAEL = 0.21
No models provided adequate fit to
the data, specifically variance
Total Ti PND 1 (litter n)b—Feng et
al. (2017)
Mouse/Fi—female
BMDLo.5sd ~ 0.095
(BMDLisd = 0.25)
Adequate model fit
Total T4 PND 30—Feng et al.
(2017)
Mouse/Fi—female
NOAEL = 0.21
No models provided adequate
statistical or visual fit to mean
responses
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Table 9. PODs Considered for Deriving the Subchronic RfD for

K+PFBS (CASRN 29420-49-3)

Endpoint/Reference
Species/Life
Stage—Sex
POD (HED)a
(mg/kg-d)
Comments*
Total T4 PND 60—Feng et al.
(2017)
Mouse/Fi—female
NOAEL = 0.21
No models provided adequate fit to
the data, specifically variance
TSH PND 30—Feng et al. (2017)
Mouse/Fi—female
NOAEL = 0.21
No models provided adequate
statistical or visual fit to mean
responses
Total T4—NTP (2019)
Rat—male
I.OAF.I. = 0.29
No models provided adequate
statistical or visual fit to mean
responses

Rat—female
BMDLisd = 0.037
Adequate model fit
Free T4—NTP (2019)
Rat—male
I.OAF.I. = 0.34
No models provided adequate
statistical or visual fit to mean
responses

Rat—female
BMDLisd = 0.027
Adequate model fit
Developmental effects
Eves opening (fetal n)h—Feng et
al. (2017)
Mouse/Fi—female
NOAEL = 0.21
No models provided adequate fit to
the data, specifically variance
Eves opening (litter n)h—Feng et
al. (2017)
Mouse/Fi—female
BMDLo.5sd = 0.073
(BMDLisd = 0.16)
Adequate model fit
Vaginal ODcning (fetal n)h—Feng
et al. (2017)
Mouse/Fi—female
BMDLo.5sd = 0.15
(BMDLisd = 0.35)
Adequate model fit
Vaginal ODcning (littern)h—Feng
et al. (2017)
Mouse/Fi—female
BMDLo.5sd = 0.094
(BMDLisd = 0.22)
Adequate model fit
First estrous (fetal n)h—Feng et al.
(2017)
Mouse/Fi—female
NOAEL = 0.21
No models provided adequate
statistical or visual fit to mean
responses
First estrous (litter n)h—Feng et al.
(2017)
Mouse/Fi—female
NOAEL = 0.21
No models provided adequate
statistical or visual fit to mean
responses
Kidney effects
Kidney histopathology—papillary
Rat—male
BMDLio = 0.49
Adequate model fit
epithelial tubular/ductal
hvDcrolasia—Lieder et al. (2009a)
Rat—female
BMDLio ~ 0.30
Adequate model fit
Kidney histopathology—papillary
Rat/Po—male
BMDLio = 0.35
Adequate model fit
epithelial tubular/ductal
hvDerolasia—Lieder et al. (2009b)
Rat/Po—female
BMDLio = 0.27
Adequate model fit
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Table 9. PODs Considered for Deriving the Subchronic RfD for
K+PFBS (CASRN 29420-49-3)
Endpoint/Reference
Species/Life
Stage—Sex
POD (HED)a
(mg/kg-d)
Comments*
Kidney histopathology—papillary
epithelial tubular/ductal
hvoerolasia—Lieder et al. (2009b)
Rat/Fi—male
BMDLio — 0.78
Adequate model fit
Rat/Fi—female
BMDLio = 0.48
Adequate model fit
'Following U.S. EPA (2011b) and U.S. EPA (2014d) guidance, animal doses from candidate principal studies were
converted to HEDs by applying a DAF, where HED = dose x DAF.
' Fetal endpoints from Feng etal. (2017) were modeled alternatively using dose-group sizes based either on total
number of fetuses or dams. Given that Feng etal. (2017) seems not to have used the litter as the statistical unit of
analysis, it is unclear whether the study-reported standard errors pertain to litters or fetuses. Alternatively,
modeling fetal endpoints using litter n or fetal n provides two modeling results that bracket the "true" variance
among all fetuses in a dose group (i.e., using the fetal n will underestimate the true variance while using the litter n
will overestimate the true variance). Individual animal data were requested from study authors but were unable to
be obtained.
BMD modeling methods and links to modeling inputs and results in HAWC are found in Appendix F. HAWC
visualization: Candidate PODs for subchronic and chronic RfD.
BMDLo 5sd = benchmark dose lower confidence limit for 0.5 SD change from the control;
BMDLio = 10% benchmark dose lower confidence limit; BMDLiSD = benchmark dose lower confidence limit for
1 SD change from the control; DAF = dosimetric adjustment factor; HAWC = Health Assessment Workspace
Collaborative; HED = human equivalent dose; K+PFBS = potassium perfluorobutane sulfonate;
LOAEL = lowest-observed-adverse-effect level; NOAEL = no-observed-adverse-effect level; PND = postnatal
day; POD = point of departure; RfD = oral reference dose; SD = standard deviation; T4 = thyroxine; TSH = thyroid
stimulating hormone.
6.1.1.3 Considerations in Selecting the Critical Effect for Deriving Oral Reference Doses
The evidence for the thyroid, developmental, and kidney effect domains support a hazard
via the oral exposure route (see Table 7). However, there are qualitative and quantitative
differences in the strength of evidence between these effect domains (see Table 9).
PFBS-induced perturbation of the thyroid was consistently observed across two species, sexes,
life stages, and exposure durations in two independent, high-confidence studies. These
perturbations involved a coherent pattern of hormonal changes with similar sensitivity in the
POD ranges across life stages (e.g., maternal and PND 1/newborn BMDLoss of 0.093 and
0.095 mg/kg-day, respectively). Developmental effects (e.g., delayed eyes opening, vaginal
opening, or first estrous) were observed in mouse litters in which decrements in thyroid hormone
occurred and with similar sensitivity in the ranges of POD estimates
(i.e., 0.073-0.21 mg/kg-day) (Feng et al.. 2017). However, these developmental effects have
been reported in a single study and species (mouse). Kidney effects in adult animals (Lieder et
al.. 2009a; Lieder et al.. 2009b) were observed in adult or developing rats across high- or
medium-confidence gavage studies of various duration; however, were less sensitive at
0.27 mg/kg-day and above.
In deriving a subchronic RfD, both the Feng et al. (2017) and NTP (2019) studies were
considered as potential principal studies because of the observed sensitivity of thyroid hormone
decrements. However, the biological significance of hypothyroxinemia (i.e., decreased T4) in
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adult euthyroid animals, absent additional signs of overt thyroid toxicity (e.g., reflex increase in
TSH and/or alterations in tissue weight or histology), is unclear; therefore, the thyroid effects
from the NTP (2019) rat study were not selected as a critical effect. The gestational exposure
study in mice was selected as the principal study for deriving the subchronic RfD based on
thyroid effects. The gestational exposure study conducted by Feng et al. (2017) reported
administration of K+PFBS by gavage in ICR mice (10/dose) from GDs 1 to 20. This study was
of good quality (i.e., high confidence) with adequate reporting and consideration for appropriate
study design, methods, and conduct (click to see risk of bias analysis in HAWC). Feng et al.
(2017) reported statistically significantly decreased total T3, total T4, and free T4, as well as
increased TSH in dams and offspring (increased TSH PND 30 only) gestationally exposed to
PFBS.
The critical effect from the Feng et al. (2017) study was decreased serum total thyroxine
(T4) in newborn (PND 1) mice. T4 and T3 are essential for normal growth of developing
offspring across animal species [for review see Forhead and Fowden (2014)1. And, previous
studies have shown that exposure to other PFAS during pregnancy results in lower T4 and T3
levels in pregnant women and fetuses or neonates (Yang et al.. 2016; Wang et al.. 2014). The
selection of total T4 as the critical effect is based on a number of key considerations (see below)
that account for cross-species correlations in thyroid physiology and hormone dynamics
particularly within the context of a developmental life stage.
A key consideration for selecting total T4 is that this represents the aggregate of potential
thyroid endocrine signaling (i.e., free T4 + protein bound T4) at any given time. In humans, FT4
represents approximately 0.03% of circulating hormone, indicating that as much as 99.97% of all
T4 is protein bound (e.g., albumin; TBG). Although T3 is the active hormone form in respondent
somatic tissues, the formation of T3 is contingent upon the deiodination of free T4. A critical
consideration in pregnant females is that T4, not T3, is the thyroid hormone that crosses the
placenta of humans and rodents. Although free T4 might be considered a suitable measure of
thyroid hormone status in nondevelopmental (e.g., adult) life stages, there are some important
factors associated with maintenance of the microenvironment for developing offspring in utero
that supports using total T4 as the critical effect. A tightly regulated transfer of maternal thyroid
hormone to a fetus is paramount to proper development of multiple tissues and organ systems
(e.g., nervous system), especially during the early trimesters. The placenta has transporters and
deiodinases that collectively act as a gatekeeper to maintain an optimal T4 microenvironment in
the fetal compartment (Fisher, 1997; Koopdonk-Kool et al.. 1996). For example, deiodinase 3
(D3) is highly expressed in human uterus, placenta, and amniotic membrane, where it serves a
critical role of regulating thyroid hormone transfer to the fetus through the deiodination of T4 to
transcriptionally inactive reverse triiodothyronine (rT3) or T3 to inactive 3,5-diiodo-L-thyronine
(T2). Similarly, Wasco et al. (2003) showed that D3 is highly expressed in the rodent uterus and
is highly induced during pregnancy. Further, the Dio3 gene that encodes D3 has been shown to
be imprinted in the mouse (Hernandez et al.. 2002), suggesting a pivotal role for this specific
deiodinase in the mouse as well. Indeed, the human and rodent placenta have been shown to be
similarly permeable to T4 and T3 (Fisher, 1997; Calvo et al.. 1992). Due to placental barrier
functionality, free T4 levels in a pregnant dam might not be entirely representative of actual T4
status in a developing fetus. Further, the American Thyroid Association published a guidelines
document in 2017 in which they stated: "Current uncertainty around FT4 estimates in pregnancy
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has led some to question the wisdom of relying on any FT4 immunoassays during pregnancy. In
contrast, measurement of TT4 and the calculated FT4 index do show the expected inverse
relationship with serum TSH. This finding suggests that TT4 measurements may be superior to
immunoassay measurement of FT4 measurements in pregnant women" (Alexander et al.. 2017).
Thus, decreased total T4 in offspring (and dams during pregnancy/at delivery) is expected to be
more representative of PFBS-mediated thyroid effects and potentially associative developmental
effects.
There are some differences in HPT development and functional maturation and
regulation during early life stages (e.g., timing of in utero and early postnatal thyroid
development) between humans and rodents [for a comprehensive overview see Regulatory
Science Associates (2019)1. Human thyroid development occurs in three phases in utero which
entails initial development of the gland between Embryonic Day 10 to Gestational Week 11
(Phase I), maturation of the fetal thyroid system from Gestational Weeks 11-35 (Phase II), and
further refinement of hypothalamic-pituitary-thyroid axis functionality during the latter portion
of gestation up to approximately 4 weeks into the postnatal period (Phase 111) (Klein et al.. 1982;
Fisher and Klein. 1981). Importantly, in utero development of the rodent thyroid gland occurs in
the same phases and order as humans, the difference being that rodents are essentially born
during Phase II, with Phase III occurring almost exclusively postnatally; whereas in humans,
Phase III is well underway in utero and completes postnatally. Accordingly, rodent
neurodevelopment in the early postnatal phase is analogous to the third trimester of human
development in utero (Gilbert et al .. 2012). Further, fetal development of rodents in utero is
entirely dependent on maternal thyroid hormone until approximately GD 17-18, whereas in
humans, fetal development transitions from complete reliance on maternal thyroid hormone
during the first trimester (i.e., thyroid development Phase I) to a mix of fetal thyroid hormone
synthesis and maternal transplacental hormone transfer beginning in the second trimester
(i.e., thyroid development Phase II) through the in utero portion of Phase 111 (Fisher and Klein.
1981).
Within the context of early developmental life stages, there are several commonalities in
HPT dynamics between humans and rodents such as similar profiles of (1) thyroid hormone
binding proteins, (2) hormone functional reserve, and (3) placental deiodinase. For example, two
carrier proteins—thyroid binding globulin (TBG) and transthyretin (TTR)—are primarily
responsible for storage and transit of T4 in mammals (Rabah et al.. 2019). TBG is the primary
carrier of T4 in humans across all life stages (Savu et al.. 1991). Importantly, in fetal and infant
rats, TBG is also the primary carrier of T4 (Savu et al.. 1989). As rats transition to adulthood,
TTR takes over as the primary carrier of T4. In addition, as a relatively highly abundant carrier
protein, albumin also plays a role in thyroid hormone binding and transit in humans and rodents;
however, the relative affinity for binding is lower than either TBG or TTR.
Life-stage-specific differences in thyroid hormone reserve capacity between adults and
neonates have been noted. On average, intrathyroidal thyroglobulin stores in adults are on the
order of months, whereas in neonates the functional reserve is approximated at less than 1 day
(Gilbert and Zoetter, 2010; Savin et al.. 2003; van den Hove et al.. 1999). This suggests that the
adult thyroid has compensatory abilities not present in early life stages, making fetal/neonatal
populations particularly sensitive to perturbations in thyroid hormone economy
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(e.g., hypothyroxinemia). And although the timing of thyroid development can vary between
species (Forhead and Fowden. 2014). the dynamic reserve capacity of T4 between humans and
rodents near birth and in early postpartum might not be significantly different. For example,
human neonates have a serum half-life of T4 of approximately 3 days (Vulsma et al.. 1989). and
thyroid tissue stores of T4 are estimated to be less than 1 day (van den Hove et al.. 1999).
Because the developing rodent thyroid does not begin producing its own hormone until late in
gestation (>GD 17), newborn rodent T4 levels are primarily a reflection of transplacental^
translocated maternal hormone; and adult rats have been shown to have a serum T4 half-life of
0.5-1 day (Choksi et al.. 2003). For this reason, significant differences in functional thyroid
reserve capacity between human and rodent neonates are not anticipated.
Accounting for the information presented above, the subchronic RfD, based on the
BMDLo 5sd (FLED) of 0.095 mg/kg-day for decreased serum total T4 in newborn (PND 1) mice,
is derived as follows:
Subchronic RfD for K+PFBS = BMDLo.ssd (HED) - UFc
= 0.095 mg/kg-day 100
= 0.00095mg/kg-day
= 1 x 10"3 mg/kg-day
Table 10 summarizes the uncertainty factors for the subchronic RfD for K+PFBS based
on effects in the thyroid.
Table 10. Uncertainty Factors for the Subchronic RfD for Thyroid Effects for
K+PFBS (CASRN 29420-49-3)
UF
Value
Justification
UFa
3
A UFa of 3 (10°5) is applied to account for uncertainty in characterizing the toxicokinetic and
toxicodynamic differences between mice and humans following oral K+PFBS/PFBS exposure. Some
aspects of the cross-species extrapolation of toxicokinetic and toxicodynamic processes have been
accounted for by calculating an HED by applying a DAF as outlined in the U.S. EPA's Recommended
Use of Bodv Weishf/4 as the Default Method in Derivation of the Oral Reference Dose ('U.S. EPA.
201 lb). However, some residual uncertainty remains in the relative cross-soecies sensitivity in
toxicodynamics (e.g., thyroid signaling). Thus, in the absence of chemical-specific data to quantify
these uncertainties, U.S. EPA's guidance recommends use of a UFA of 3.
UFd
3
A UFd of 3 is applied due to database deficiencies. The oral exposure database contains multiple
short-term and subchronic-duration toxicity studies of laboratory animals (NTP. 2019; Biiland et al..
2011; 3M. 2010; Lieder et al.. 2009a; 3M. 2001. 2000d). a two-generation reproductive toxicity study
in rats (Lieder et al.. 2009b). and multiple developmental toxicity studies in mice and rats (Feng et al..
2017; York. 2002). However, the observation of decreased thvroid hormone is known to be a crucial
element during developmental life stages, particularly for neurodevelopment, and the database is
limited by the lack of developmental neurotoxicity studies. In addition, because other health effect
domains such as immunotoxicity and mammary gland development are effects of increasing concern
across several members of the larger PFAS family (Grandiean. 2018; Liew et al.. 2018; White et al..
2007). the lack of studies evaluating these outcomes following PFBS exposure is a limitation in the
database.
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Table 10. Uncertainty Factors for the Subchronic RfD for Thyroid Effects for
K+PFBS (CASRN 29420-49-3)
UF
Value
Justification
UFh
10
A UFh of 10 is applied to account for interindividual variability in the human populations because of
both intrinsic (toxicokinetic, toxicodynamic, genetic, life stage, and health status) and extrinsic (life
style) factors that can influence the response to dose. In the absence of chemical-specific data to
quantify this variability in the toxicokinetics and toxicodynamics of K+PFBS/PFBS in humans,
U.S. EPA recommends using a UFH of 10.
UFl
1
A UFl of 1 is applied for LOAEL-to-NOAEL extrapolation because the POD is a BMDL and the
BMR was selected based on evidence that it represented a minimal biologically significant response
level in susceptible populations such as developing offspring.
UFS
1
A UFS of 1 is applied because the POD comes from a developmental study in mice. The
developmental period is recognized as a susceptible life stage in which exposure during certain time
windows (e.g., gestational) is more relevant to the induction of developmental effects than lifetime
exposure (U.S. EPA. 1991a).
UFC
100
Composite UF = UFA x UFD x UFH x UFL x UFS
BMDL = benchmark dose lower confidence limit; BMR = benchmark response; DAF = dosimetric adjustment
factor; HED = human equivalent dose; K+PFBS = potassium perfluorobutane sulfonate;
LOAEL = lowest-observed-adverse-effect level; NOAEL = no-observed-adverse-effect level; PFAS = per- and
polyfluoroalkyl substances; PFBS = perfluorobutane sulfonic acid; POD = point of departure; RfD = oral reference
dose; UF = uncertainty factor; UFA = interspecies uncertainty factor; UFC = composite uncertainty factor;
UFd = database uncertainty factor; UFH = intraspecies uncertainty factor; UFL = LOAEL-to-NOAEL uncertainty
factor; UFS = subchronic-to-chronic uncertainty factor.
The data for K+PFBS can be used to derive a subchronic RfD for the free acid (PFBS), as
K PFBS is fully dissociated in water at the environmental pH range of 4-9 (NICNAS. 2005). To
calculate the subchronic RfD for the free acid, the subchronic RfD for the potassium salt is
adjusted to compensate for differences in MW between K+PFBS (338.19) and PFBS (300.10).
The subchronic RfD for PFBS (free acid) is calculated as follows:
Subchronic RfD	=	RfD for K+PFBS salt x (MW free acid ^ MW salt)
for PFBS (free acid) =	0.00095 mg/kg-day x (300.10 - 338.19)
=	0.00095 mg/kg-day x (0.89)
=	0.00085 mg/kg-day
=	9 x 10"4 mg/kg-day
Confidence in the subchronic RfD for PFBS and K+PFBS for thyroid effects is medium,
as explained in Table 11.
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Table 11. Confidence Descriptors for the Subchronic RfD for PFBS (CASRN 375-73-5)
and the Related Compound K+PFBS (CASRN 29420-49-3)
Confidence Categories
Designation
Discussion
Confidence in study
H
Confidence in the principal study is high because the overall
study design, performance, and characterization of exposure
was sood. Studv details and risk of bias analysis can be found
in HAWC.
Confidence in database
M
Confidence in the oral toxicity database for derivation of the
candidate subchronic RfD for thyroid effects is medium
because although there are multiple developmental toxicity
studies in mice and rats, no studies are available that have
specifically evaluated neurodevelopmental, immunological, or
mammary gland effects. In addition, available toxicokinetic
studies are limited (e.g., one mouse toxicokinetic study) and
toxicokinetic data do not exist for PFBS at all life stages,
including neonates, infants, and children. Additionally,
studies are not available to estimate the relative cross-species
sensitivity in toxicodynamics (e.g., thyroid signaling).
Confidence in candidate subchronic
RfD
M
The overall confidence in the candidate subchronic RfD for
thyroid effects is medium.
H = high; HAWC = Health Assessment Workspace Collaborative; K+PFBS = potassium perfluorobutane sulfonate;
M = medium; PFBS = perfluorobutane sulfonic acid; RfD = oral reference dose.
The subchronic RfD is derived to be protective of all types of effects across studies and
species following oral subchronic exposure and is intended to protect sensitive subpopulations
and life stages.
6.1.2 Derivation of the Chronic Oral Reference Dose
There are no chronic studies available for PFBS and K+PFBS. Therefore, based on the
same database and similar considerations as the subchronic RfD, the noncancer chronic RfD is
derived, based on the same BMDLo.ssd (HED) of 0.095 mg/kg-day for decreased serum total T4
in newborn (PND 1) mice (Feng et al.. 2017), as follows:
Chronic RfD for K+PFBS = BMDLo.ssd (HED) - UFc
= 0.095 mg/kg-day 300
= 0.00032 mg/kg-day
= 3 x 10"4 mg/kg-day
Table 12 summarizes the uncertainty factors for the chronic RfD for K+PFBS based on
effects in the thyroid.
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Table 12. Uncertainty Factors for the Chronic RfD for Thyroid for
K+PFBS (CASRN 29420-49-3)
UF
Value
Justification
UFa
3
A UFa of 3 (10°5) is applied to account for uncertainty in characterizing the toxicokinetic and
toxicodynamic differences between mice and humans following oral K+PFBS/PFBS exposure. Some
aspects of the cross-species extrapolation of toxicokinetic and toxicodynamic processes have been
accounted for by calculating an HED by applying a DAF as outlined in the U.S. EPA's Recommended
Use of Bodv Weight3/4 as the Default Method in Derivation of the Oral Reference Dose (U.S. EPA.
201 lb). However, some residual uncertainty remains in the relative cross-species sensitivity in
toxicodynamics (e.g., thyroid signaling). Thus, in the absence of chemical-specific data to quantify
these uncertainties, U.S. EPA's guidance recommends using a UFA of 3.
UFd
10
A UFd of 10 is applied to account for database deficiencies. The oral exposure database contains
multiple short-term and subchronic-duration toxicity studies of laboratory animals (NTP. 2019;
Biiland et al. 2011; Lieder et al. 2009a: 3M. 2001. 2000d). a two-generation reproductive toxicity
study in rats (Lieder et al.. 2009b). and multiple developmental toxicity studies in mice and rats (Feng
et al.. 2017; York. 2002). However, because thyroid hormone is known to be critical during
developmental life stages, particularly for neurodevelopment, the database is limited by the lack of
developmental neurotoxicity studies. Further, because of the lack of chronic studies, there is
additional uncertainty regarding how longer-term exposures might affect hazard identification and
dose-response assessment for PFBS via the oral route (e.g., potentially more sensitive effects).
Lastly, because immunotoxicity and mammary gland development are effects of increasing concern
across several members of the larger PFAS family (Grandiean. 2018; Liew et al.. 2018; White et al..
2007). the lack of studies evaluating these outcomes following PFBS exposure is a limitation in the
database.
UFh
10
A UFh of 10 is applied to account for interindividual variability in the human populations because of
both intrinsic (toxicokinetic, toxicodynamic, genetic, life stage, and health status) and extrinsic
(lifestyle) factors that can influence the response to dose. In the absence of chemical-specific data to
quantify this variability in the toxicokinetics and toxicodynamics of K+PFBS/PFBS in humans,
U.S. EPA recommends using a UFH of 10.
UFl
1
A UFl of 1 is applied for LOAEL-to-NOAEL extrapolation because the POD is a BMDL and the
BMR was selected based on evidence that it represented a minimal biologically significant response
level in susceptible populations such as developing offspring.
UFS
1
A UFS of 1 is applied because the POD comes from a developmental study of mice. The
developmental period is recognized as a susceptible life stage in which exposure during certain time
windows (e.g., gestational) is more relevant to the induction of developmental effects than lifetime
exposure (U.S. EPA. 1991b). The additional concern over potential hazards following longer term
(chronic) exposures is accounted for under the UFD above.
UFC
300
Composite UF = UFA x UFD x UFH x UFL x UFS
BMDL = benchmark dose lower confidence limit; BMR = benchmark response; DAF = dosimetric adjustment
factor; HED = human equivalent dose; K+PFBS = potassium perfluorobutane sulfonate;
LOAEL = lowest-observed-adverse-effect level; NOAEL = no-observed-adverse-effect level; PFAS = per- and
polyfluoroalkyl substances; PFBS = perfluorobutane sulfonic acid; POD = point of departure; RfD = oral reference
dose; UF = uncertainty factor; UFA = interspecies uncertainty factor; UFC = composite uncertainty factor;
UFd = database uncertainty factor; UFH = intraspecies uncertainty factor; UFL = LOAEL-to-NOAEL uncertainty
factor; UFS = subchronic-to-chronic uncertainty factor.
The data for K+PFBS can be used to derive a chronic RfD for the free acid (PFBS),
because K PFBS is fully dissociated in water at the environmental pH range of 4-9 (NICNAS.
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2005). To calculate the chronic RfD for the free acid, the chronic RfD for the potassium salt is
adjusted to compensate for differences in MW between K+PFBS (338.19) and PFBS (300.10).
The chronic RfD for PFBS (free acid) for thyroid effects is the same as the value for the K+PFBS
salt. The chronic RfD for PFBS (free acid) is calculated as follows:
Chronic RfD	=	RfD for K+PFBS salt x (MW free acid - MW salt)
for PFBS (free acid) =	0.00032 mg/kg-day x (300.10 - 338.19)
=	0.00032 mg/kg-day x (0.89)
=	0.00028 mg/kg-day
=	3 x 10"4 mg/kg-day
Confidence in the chronic RfD for PFBS and K+PFBS for thyroid effects is low, as
explained in Table 13 below.
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Table 13. Confidence Descriptors for Chronic RfD for PFBS (CASRN 375-73-5) and the
Related Compound K+PFBS (CASRN 29420-49-3)
Confidence Categories
Designation
Discussion
Confidence in study
H
Confidence in the principal study is high because the overall study
design, performance, and characterization of exposure was good.
Studv details and risk of bias analysis can be found in HAWC.
Confidence in database
L
Confidence in the oral toxicity database for deriving the chronic RfD is
low because, although there are multiple short-term studies and a
subchronic-duration toxicity study in laboratory animals, one
acceptable two-generation reproductive toxicity study in rats, and
multiple developmental toxicity studies in mice and rats, the database
lacks any chronic-duration exposure studies or studies that have
evaluated neurodevelopmental, immunological, or mammary gland
effects. In addition, available toxicokinetic studies are limited
(e.g., one mouse toxicokinetic study) and toxicokinetic data do not
exist for PFBS at all life stages, including neonates, infants, and
children. Additionally, studies are not available to estimate the relative
cross-species sensitivity in toxicodynamics (e.g., thyroid signaling).
Confidence in candidate
chronic RfD
L
The overall confidence in the candidate chronic RfD for thyroid effects
is low.
H = high; HAWC = Health Assessment Workspace Collaborative; K+PFBS = potassium perfluorobutane sulfonate;
L = low; PFBS = perfluorobutane sulfonic acid; RfD = oral reference dose.
The chronic RfD is derived to be protective of all types of effects across studies and
species following oral chronic exposure and is intended to protect the population as a whole,
including potentially susceptible populations and life stages (U.S. EPA. 2002). This value
should be applied in general population risk assessments. Decisions concerning averaging
exposures over time for comparison with the RfD should consider the types of toxicological
effects and specific life stages of concern. For example, fluctuations in exposure levels that
result in elevated exposures during development could potentially lead to an appreciable risk,
even if average levels over the full exposure duration were less than or equal to the RfD.
6.2 DERIVATION OF INHALATION REFERENCE CONCENTRATIONS
No published studies investigating the effects of subchronic- or chronic-duration
inhalation toxicity of PFBS and the related compound K+PFBS in humans or animals have been
identified.
6.3 CANCER WEIGHT-OF-EVIDENCE DESCRIPTOR AND DERIVATION OF
CANCER RISK VALUES
No studies evaluating the carcinogenicity of PFBS or K+PFBS in humans or animals have
been identified. In accordance with the Guidelines for Carcinogen Risk Assessment (U.S. EPA.
2005). the U.S. EPA concluded that there is "Inadequate Information to Assess Carcinogenic
Potential" for PFBS and K+PFBS by any route of exposure. Therefore, the lack of data on the
carcinogenicity of PFBS and the related compound K+PFBS precludes the derivation of
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quantitative estimates for either oral (oral slope factor) or inhalation (inhalation unit risk)
exposure.
6.4 SUSCEPTIBLE POPULATIONS AND LIFE STAGES
Early life stages as well as pregnant women are potentially susceptible to PFBS exposure.
PFBS has been detected in blood serum of nursing mothers, which might indicate a potential for
lactational exposure (Glynn et al.. 2012); however, information on the kinetics of lactational
transfer are lacking and represents a key data gap for future research.
The available information suggests sex-specific variation in the toxicokinetics of PFBS in
rodents. Studies in mice and rats generally report clearance and elimination half-lives to be
faster for females than for males (see the "Toxicokinetics" section). For example, Lau et al.
(2020) reports statistically significant differences in half-life between the sexes with female mice
exhibiting a shorter half-life than males. Similar sex-specific variation in elimination has been
reported in rats. 01 sen et al. (2009) reported a statistically significant difference in the urinary
clearance rates (p < 0.01), with female rats (469 ± 40 mL/hour) having faster clearance rates than
male rats (119 ± 34 mL/hour). Huang et al. (2019a) also reported higher clearance in female rats
than in male rats given the same dose (26.0-75.5 mL/hour-kg in males, 152-259 mL/hour-kg in
females). Chengelis et al. (2009) reported that the mean apparent clearance of PFBS from the
serum was approximately eightfold higher for female rats (0.311 L/hour-kg) than for male rats
(0.0394 L/hour-kg). Statistically significant sex-related differences in half-life or clearance were
not observed between male and female monkeys (01 sen et al.. 2009). Differences in the
toxicokinetics in rodents could result in sex-specific differences in toxicity studies.
In vivo toxicity studies report that PFBS exposure can alter thyroid hormone levels in
parental and Fi generation animals (see the "Thyroid Effects" section). Thyroid hormones play a
critical role in coordinating complex developmental processes for various organs/systems
(e.g., reproductive and nervous system), and disruption of thyroid hormone production/levels in a
pregnant woman or neonate can have persistent adverse health effects for the developing
offspring (Ghassabian and Trasande. 2018; Foster and Gray. 2013; Julvez et al.. 2013; Roman et
al.. 2013).
Animal studies also provide evidence that gestationally exposed females might be a
susceptible subpopulation because of potential effects on female reproduction, including
evidence of altered ovarian follicle development and delayed vaginal opening (see the
"Reproductive Effects" section). Furthermore, gestationally exposed females also had
significantly reduced BWs and delayed eye opening. These findings suggest that developmental
landmarks indicative of adverse responses can be affected after PFBS exposure (see the
"Offspring Growth and Early Development" section).
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APPENDIX A. LITERATURE SEARCH STRATEGY
This appendix presents the full details of the literature search strategy used to identify
primary, peer-reviewed literature pertaining to perfluorobutane sulfonic acid (PFBS) (Chemical
Abstracts Service registry number [CASRN] 375-73-5) and/or the potassium salt (K+PFBS)
(CASRN 29420-49-3) and the deprotonated anionic form of PFBS (i.e., PFBS"; CASRN
45187-15-3). Initial database searches were conducted on July 18, 2017 using four online
scientific databases (PubMed, Web of Science [WOS], TOXLINE, and TSCATS via TOXLINE)
and updated on February 28, 2018; May 1, 2019; and May 15, 2020. The literature search
focused on chemical name and synonyms (see Table A-l) with no limitations on publication
type, evidence stream (i.e., human, animal, in vitro, and in silico) or health outcomes. Beyond
database searches, references were also identified from studies submitted under the Toxic
Substances Control Act (TSCA) and from review of other government documents (e.g., Agency
for Toxic Substances and Disease Registry [ATSDR]) and combined with the results of the
database search. Search results are retained in the U.S. EPA's Health and Environmental
Research Online (HERO) database.
Table A-l. Synonyms and MeSH Terms
ChemID
375-73-5
1,1,2,2,3,3,4,4,4-Nonafluoro-l-butanesulfonic acid
1-Perfluorobutanesulfonic acid
Nonafluoro-l-butanesulfonic acid
Nonafluorobutanesulfonic acid
Perfluorobutanesulfonic acid
PFBS
1,1,2,2,3,3,4,4,4-Nonafluorobutane-l-sulphonic acid
PubMed (new only)
Perfluorobutane sulfonic acid
Perfluorobutanesulfonate
Perfluorobutane sulfonate
EPA Spreadsheet
1,1,2,2,3,3,4,4,4-Nonafluoro-l-butanesulfonic acid
1-Butanesulfonic acid, 1,1,2,2,3,3,4,4,4-nonafluoro-
1-Butanesulfonic acid, nonafluoro-
1-Perfluorobutanesulfonic acid
Nonafluoro-l-butanesulfonic acid
Nonafluorobutanesulfonic acid
PFBS
Perfluoro-1 -butanesulfonate
Perfluorobutane sulfonate
Perfluorobutanesulfonate
Perfluorobutanesulfonic acid
Perfluorobutylsulfonate
45187-15-3
MeSH = medical subject headings; PFBS = perfluorobutane sulfonic acid.
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A.l. LITERATURE SEARCH STRINGS
PubMed
375-73-5[rn] OR 45187-15-3[rn] "nonafluorobutane-l-sulfonic acid"[nm] OR
"1,1,2,2,3,3,4,4,4-Nonafluoro-l-butanesulfonic acid"[tw] OR "1-Perfluorobutanesulfonic
acid"[tw] OR "Nonafluoro-l-butanesulfonic acid"[tw] OR "Nonafluorobutanesulfonic acid"[tw]
OR "Perfluorobutanesulfonic acid"[tw] OR "1,1,2,2,3,3,4,4,4-Nonafluorobutane-l-sulphonic
acid"[tw] OR "Perfluorobutane sulfonic acid"[tw] OR "Perfluorobutanesulfonate"[tw] OR
"Perfluorobutane sulfonate"[tw] OR "1-Butanesulfonic acid, l,l,2,2,3,3,4,4,4-nonafluoro-"[tw]
OR "1-Butanesulfonic acid, nonafluoro-"[tw] OR "Perfluoro-l-butanesulfonate"[tw] OR
"Perfluorobutylsulfonate"[tw] OR "Eftop FBSA"[tw] OR (PFBS[tw] AND (fluorocarbon*[tw]
OR fluorotelomer*[tw] OR polyfluoro* [tw] OR perfluoro-* [tw] OR perfluoroa*[tw] OR
perfluorob*[tw] OR perfluoroc*[tw] OR perfluorod*[tw] OR perfluoroe*[tw] OR
perfluoroh*[tw] OR perfluoron*[tw] OR perfluoroo*[tw] OR perfluorop*[tw] OR
perfluoros*[tw] OR perfluorou*[tw] OR perfluorinated[tw] OR fluorinated[tw] ORPFAS[tw]
OR PFOS[tw] OR PFOA[tw]))
wos
TS="l,l,2,2,3,3,4,4,4-Nonafluoro-l-butanesulfonic acid" OR
TS=" 1-Perfluorobutanesulfonic acid" OR TS="Nonafluoro-l-butanesulfonic acid" OR
TS="Nonafluorobutanesulfonic acid" OR TS="Perfluorobutanesulfonic acid" OR
TS="l,l,2,2,3,3,4,4,4-Nonafluorobutane-l-sulphonic acid" OR TS="Perfluorobutane sulfonic
acid" OR TS="Perfluorobutanesulfonate" OR TS="Perfluorobutane sulfonate" OR
TS=" 1-Butanesulfonic acid, 1,1,2,2,3,3,4,4,4-nonafluoro-" OR TS=" 1-Butanesulfonic acid,
nonafluoro-" OR TS="Perfluoro-l-butanesulfonate" OR TS="Perfluorobutylsulfonate" OR
TS="Eftop FBSA" OR (TS=PFBS AND TS=(fluorocarbon* OR fluorotelomer* OR polyfluoro*
OR perfluoro-* OR perfluoroa* OR perfluorob* OR perfluoroc* OR perfluorod* OR
perfluoroe* OR perfluoroh* OR perfluoron* OR perfluoroo* OR perfluorop* OR perfluoros*
OR perfluorou* OR perfluorinated OR fluorinated OR PFAS OR PFOS OR PFOA))
TOXLINE
( ( 375-73-5 [rn] OR 45187-15-3 [rn] OR "1 1223344 4-nonafluoro-l-butanesulfonic
acid" OR "1-perfluorobutanesulfonic acid" OR "nonafluoro-l-butanesulfonic acid" OR
"nonafluorobutanesulfonic acid" OR "perfluorobutanesulfonic acid" OR "1 1223344
4-nonafluorobutane-l-sulphonic acid" OR "perfluorobutane sulfonic acid" OR
"perfluorobutanesulfonate" OR "perfluorobutane sulfonate" OR "1-butanesulfonic acid 112 2 3
3 4 4 4-nonafluoro-" OR "1-butanesulfonic acid nonafluoro-" OR "perfluoro-1-butanesulfonate"
OR "perfluorobutylsulfonate" OR "eftop fbsa" OR (pfbs AND (fluorocarbon* OR
fluorotelomer* OR polyfluoro* OR perfluoro* OR perfluorinated OR fluorinated OR pfas OR
pfos OR pfoa ) ) ) ) AND ( ANEUPL [org] OR BIOSIS [org] OR CIS [org] OR DART [org] OR
EMIC [org] OR EPIDEM [org] OR HEEP [org] OR HMTC [org] OR IPA [org] OR RISKLINE
[org] ORMTGABS [org] ORNIOSH [org] ORNTIS [org] ORPESTAB [org] ORPPBIB [org]
) AND NOT PubMed [org] AND NOT pubdart [org]
TSCATS
375-73-5[rn] AND tscats[org]; 45187-15-3[rn] AND tscats[org]
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APPENDIX B. DETAILED PECO CRITERIA
Table B-l. Population, Exposure, Comparator, and Outcome Criteria
PECO
Element
Evidence
Population
Human: Any population (occupational; general population including children, pregnant women, and
other sensitive populations). The following study designs will be considered most informative:
controlled exposure, cohort, case-control, or cross-sectional. Note: Case reports and case series are
not the primary focus of this assessment and will be tracked as supplemental material during the study
screening process.
Animal: Nonhuman mammalian animal species (whole organism) of any life stage (including
preconception, inutero, lactation, peripubertal, and adult stages).
In vitro models of genotoxicity: The studies will be considered PECO relevant. All other in vitro
studies will be tagged as "non-PECO relevant, but supplemental material."
Nonmammalian model systems/in vitro/in silico NOT related to genotoxicity: Nonmammalian
model systems (e.g., fish, amphibians, birds, and Caenorhabditis elegans); studies of human or
animal cells, tissues, or biochemical reactions (e.g., ligand binding assays) with in vitro exposure
regimens; bioinformatics pathways of disease analysis; and/or high throughput screening data. These
studies will be classified as non-PECO relevant, but have supplemental information.
Exposure
Human: Studies providing qualitative or quantitative estimates of exposure based on administered
dose or concentration, biomonitoring data (e.g., urine, blood, or other specimens), environmental or
occupational-setting measures (e.g., water levels or air concentrations), residential location, job title
or other relevant occupational information. Human "mixture" studies are considered PECO relevant
as long as they have the PFAS of interest.
Animal: Studies providing qualitative and quantitative estimates of exposure based on administered
dose or concentration. Oral and inhalation studies are considered PECO relevant. Nonoral and
noninhalation studies are tagged as supplemental. Experimental mixture studies are included as
PECO relevant only if they include a PFBS-only arm. Otherwise, mixture studies are tagged as
supplemental.
All studies must include exposure to PFBS, CASRN 375-73-5. Studies of precursor PFAS that
identify any of the targeted PFAS as metabolites will also be included.
Comparator
Human: A comparison or reference population exposed to lower levels (or no exposure/exposure
below detection levels) or for shorter periods of time. For D-R purposes, exposure-response
quantitative results must be presented in sufficient detail such as regression coefficients presented
with statistical measure of variation such as RR, HR, OR, or SMR or observed cases vs. expected
cases (common in occupational studies); slope or linear regression coefficient (i.e., per unit increase
in a continuous outcome); difference in the means; or report means with results of t-test, mean
comparison by regression, or other mean-comparing hypothesis test.
Animal: Quantitative exposure versus lower or no exposure with concurrent vehicle control group.
Outcome
Cancer and noncancer health outcomes. In general, endpoints related to clinical diagnostic criteria,
disease outcomes, histopathological examination, genotoxicity, or other apical/phenotypic outcomes
will be prioritized for evidence synthesis. Based on preliminary screening work and other
assessments, the systematic review is anticipated to focus on liver (including serum lipids),
developmental, reproductive, neurological, developmental neurotoxicity, thyroid disease/disruption,
immunological, cardiovascular, and musculoskeletal outcomes.
D-R = dose-response; HR = hazard ratio; OR = odds ratio; PECO = Population, Exposure, Comparator, and
Outcome; PFAS = per- and polyfluoroalkyl substances; PFBS = perfluorobutane sulfonic acid; RR = risk ratio;
SMR = standardized mortality ratio
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APPENDIX C. STUDY EVALUATION METHODS
For each outcome in a study, in each domain, reviewers reached a consensus judgment of
good, adequate, deficient, not reported, or critically deficient. Questions used to guide the
development of criteria for each domain in epidemiology studies are presented in Table C-l and
experimental animal toxicology studies in Table C-3. These categories were applied to each
evaluation domain for each study as follows:
• Good represents a judgment that the study was conducted appropriately in relation to the
evaluation domain and any deficiencies, if present, are minor and would not be expected
to influence the study results.
• Adequate indicates a judgment that there are methodological limitations relating to the
evaluation domain, but that those limitations are not likely to be severe or to have a
notable impact on the results.
• Deficient denotes identified biases or deficiencies that are interpreted as likely to have
had a notable impact on the results or that prevent interpretation of the study findings.
• Not reported indicates that the information necessary to evaluate the domain was not
available in the study. Generally, this term carries the same functional interpretation as
deficient for the purposes of the study confidence classification. Depending on the
number and severity of other limitations identified in the study, it may or may not be
worth reaching out to the study authors for this information.
• Critically deficient reflects a judgment that the study conduct introduced a serious flaw
that makes the observed effect(s) uninterpretable. Studies with a determination of
critically deficient in an evaluation domain will almost always cause the study to be
considered overall uninformative.
Once the evaluation domains were rated, the identified strengths and limitations were
considered to reach a study confidence rating of high, medium, low, or uninformative for a
specific health outcome. This was based on the reviewer judgments across the evaluation
domains and included consideration of the likely impact the noted deficiencies in bias and
sensitivity, or inadequate reporting, have on the results. The ratings, which reflect a consensus
judgment between reviewers, are defined as follows:
• High. A well-conducted study with no notable deficiencies or concerns were identified;
the potential for bias is unlikely or minimal, and the study used sensitive methodology.
High confidence studies generally reflect judgments of good across all or most evaluation
domains.
• Medium: A satisfactory (acceptable) study in which deficiencies or concerns were noted,
but the limitations are unlikely to be of a notable degree. Generally, medium confidence
studies will include adequate or good judgments across most domains, with the impact of
any identified limitation not being judged as severe.
• Low. A substandard study in which deficiencies or concerns were noted, and the potential
for bias or inadequate sensitivity could have a significant impact on the study results or
their interpretation. Typically, low confidence studies would have a deficient evaluation
for one or more domains, although some medium confidence studies could have a
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deficient rating in domain(s) considered to have less influence on the magnitude or
direction of effect estimates. Generally, low confidence results are given less weight than
high or medium confidence results during evidence synthesis and integration and are
generally not used as the primary sources of information for hazard identification or
derivation of toxicity values unless they are the only studies available. Studies rated as
low confidence only because of sensitivity concerns about bias towards the null require
additional consideration during evidence synthesis. Observing an effect in these studies
could increase confidence, assuming the study was otherwise well-conducted.
• Uninformative: An unacceptable study in which serious flaw(s) make the study results
unusable for informing hazard identification. Studies with critically deficient j udgments
in any evaluation domain will almost always be classified as uninformative (see
explanation above). Studies with multiple deficient judgments across domains might also
be considered uninformative. Uninformative studies will not be considered further in the
synthesis and integration of evidence for hazard identification or dose-response but might
be used to highlight possible research gaps.
Table C-l. Questions Used to Guide the Development of Criteria for Each Domain in
Epidemiology Studies
Core Question
Prompting Questions
Follow-Up Questions
ExDOSure
For all:
•	Does the exposure measure capture the variability in exposure
among the participants, considering intensity, frequency, and
duration of exposure?
•	Does the exposure measure reflect a relevant time window? If
not, can the relationship between measures in this time and the
relevant time window be estimated reliably?
•	Was the exposure measurement likely to be affected by a
knowledge of the outcome?
•	Was the exposure measurement likely to be affected by the
presence of the outcome (i.e., reverse causality)?
For case-control studies of occupational exposures:
•	Is exposure based on a comprehensive job history describing
tasks, setting, time period, and use of specific materials?
For biomarkers of exposure, general population:
•	Is a standard assay used? What are the intra- and interassay
coefficients of variation? Is the assay likely to be affected by
contamination? Are values less than the limit of detection dealt
with adequately?
What exposure time period is reflected by the biomarker? If the
half-life is short, what is the correlation between serial measurements
of exposure?
Is the degree of
exposure
misclassification likely
to vary by exposure
level?
If the correlation
between exposure
measurements is
moderate, is there an
adequate statistical
approach to ameliorate
variability in
measurements?
If there is a concern
about the potential for
bias, what is the
predicted direction or
distortion of the bias on
the effect estimate (if
there is enough
information)?
measurement
Does the
exposure
measure reliably
distinguish
between levels
of exposure in a
time window
considered most
relevant for a
causal effect
with respect to
the development
of the outcome?
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Table C-l. Questions Used to Guide the Development of Criteria for Each Domain in
Epidemiology Studies
Core Question
Prompting Questions
Follow-Up Questions
Outcome
For all:
•	Is outcome ascertainment likely to be affected by knowledge of,
or presence of, exposure (e.g., consider access to health care, if
based on self-reported history of diagnosis)?
For case-control studies:
•	Is the comparison group without the outcome (e.g., controls in a
case-control study) based on objective criteria with little or no
likelihood of inclusion of people with the disease?
For mortality measures:
•	How well does cause of death data reflect occurrence of the
disease in an individual? How well do mortality data reflect
incidence of the disease?
For diagnosis of disease measures:
•	Is diagnosis based on standard clinical criteria? If based on
self-report of diagnosis, what is the validity of this measure?
For laboratory-based measures (e.g., hormone levels):
•	Is a standard assay used? Does the assay have an acceptable
level of interassay variability? Is the sensitivity of the assay
appropriate for the outcome measure in this study population?
Is there a concern that
any outcome
misclassification is
nondifferential,
differential, or both?
What is the predicted
direction or distortion of
the bias on the effect
estimate (if there is
enough information)?
ascertainment
Does the
outcome
measure reliably
distinguish the
presence or
absence (or
degree of
severity) of the
outcome?
ParticiDant
For longitudinal cohort:
•	Did participants volunteer for the cohort based on knowledge of
exposure and/or preclinical disease symptoms? Was entry into
the cohort or continuation in the cohort related to exposure and
outcome?
For occupational cohort:
•	Did entry into the cohort begin with the start of the exposure?
•	Was follow-up or outcome assessment incomplete, and if so, was
follow-up related to both exposure and outcome status?
•	Could exposure produce symptoms that would result in a change
in work assignment/work status ("healthy worker survivor
effect")?
For case-control study:
•	Were controls representative of population and time periods from
which cases were drawn?
•	Are hospital controls selected from a group whose reason for
admission is independent of exposure?
•	Could recruitment strategies, eligibility criteria, or participation
rates result in differential participation relating to both disease
and exposure?
For population-based survey:
•	Was recruitment based on advertisement to people with
knowledge of exposure, outcome, and hypothesis?
Were differences in
participant enrollment
and follow-up evaluated
to assess bias?
If there is a concern
about the potential for
bias, what is the
predicted direction or
distortion of the bias on
the effect estimate (if
there is enough
information)?
Were appropriate
analyses performed to
address changing
exposures over time in
relation to symptoms?
Is there a comparison of
participants and
nonparticipants to
address whether
differential selection is
likely?
selection
Is there
evidence that
selection into or
out of the study
(or analysis
sample) was
jointly related to
exposure and to
outcome?
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Table C-l. Questions Used to Guide the Development of Criteria for Each Domain in
Epidemiology Studies
Core Question
Prompting Questions
Follow-Up Questions
Confounding
Is confounding adequately addressed by considerations in...
a.	... participant selection (matching or restriction)?
b.	... accurate information on potential confounders and statistical
adjustment procedures?
c.	... lack of association between confounder and outcome or
confounder and exposure in the study?
d.	... information from other sources?
Is the assessment of confounders based on a thoughtful review of
published literature, potential relationships (e.g., as can be gained
through directed acyclic graphing), minimizing potential overcontrol
(e.g., inclusion of a variable on the pathway between exposure and
outcome)?
If there is a concern
about the potential for
bias, what is the
predicted direction or
distortion of the bias on
the effect estimate (if
there is enough
information)?
Is confounding
of the effect of
the exposure
likely?
Analysis
•	Are missing outcome, exposure, and covariate data recognized
and, if necessary, accounted for in the analysis?
•	Does the analysis appropriately consider variable distributions
and modeling assumptions?
•	Does the analysis appropriately consider subgroups of interest
(e.g., based on variability in exposure level, duration, or
susceptibility)?
•	Is an appropriate analysis used for the study design?
•	Is effect modification considered, based on considerations
developed a priori?
•	Does the study include additional analyses addressing potential
biases or limitations (i.e., sensitivity analyses)?
If there is a concern
about the potential for
bias, what is the
predicted direction or
distortion of the bias on
the effect estimate (if
there is enough
information)?
Do the analysis
strategy and
presentation
convey the
necessary
familiarity with
the data and
assumptions?
Sensitivitv
•	Is the exposure range adequate?
•	Was the appropriate population included?
•	Was the length of follow-up adequate? Is the time/age of
outcome ascertainment optimal given the interval of exposure
and the health outcome?
•	Are there other aspects related to risk of bias or otherwise that
raise concerns about sensitivity?

Is there a
concern that
sensitivity of the
study is not
adequate to
detect an effect?
Selective
•	Are the results needed for the IRIS analysis presented (based on
a priori specification)? If not, can these results be obtained?
•	Are only statistically significant results presented?

reporting
Is there reason
to be concerned
about selective
reporting?
IRIS = Integrated Risk Information System.
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C.l. EXPOSURE MEASUREMENT EVALUATION CRITERIA
The criteria used to evaluate exposure measurement for PFBS (Table C-2) are adapted
from the criteria developed by the National Toxicology Program (NTP) Office of Health
Assessment and Translation for their assessment of the association between perfluorooctane
sulfonic acid (PFOS) and perfluorooctanoic acid (PFOA) and immune effects (NTP. 2016. 2015)
and were established prior to beginning study evaluation. Standard analytical methods for
evaluating individual per- and polyfluoroalkyl substances (PFAS) in serum or whole-blood using
quantitative techniques such as liquid chromatography-triple quadrupole mass spectrometry are
preferred (CDC. 2018; U.S. EPA. 2014b. e; ATSDR. 2009; CDC. 2009). The estimated serum
half-life of PFBS is approximately 1 month (l.au. 2015; 01 sen et at.. 2009). so unlike for some
other PFAS with longer half-lives, current exposure might not be indicative of past exposures.
Little data is available on repeated measures of PFBS in humans over time, so the reliability of a
single measure is unclear. The timing of the exposure measurement is considered in relation to
the etiologic window for each outcome being reviewed.
Table C-2. Criteria for Evaluation of Exposure Measurement in Epidemiology Studies
Exposure
Measurement Rating
Criteria
Good
All of the following:
• Evidence that exposure was consistently assessed using well-established methods
that directly measure exposure (e.g., measurement of PFAS in blood, serum, or
plasma).
• Exposure was assessed in a relevant time window for development of the outcome
(i.e., temporality is established, and sufficient latency occurred prior to disease
onset).
• There is evidence that a sufficient proportion of the exposure data measurements are
above the limit of quantification for the assay so that different exposure groups can
be distinguished based on the analyses conducted.
• The laboratory analysis included standard quality control measures with
demonstrated precision and accuracy.
• There is sufficient specificity/sensitivity and range or variation in exposure
measurements that would minimize potential for exposure measurement error and
misclassificationby allowing exposure classifications to be differentiated (i.e., can
reliably categorize participants into groups such as high vs. low exposure).
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Table C-2. Criteria for Evaluation of Exposure Measurement in Epidemiology Studies
Exposure
Measurement Rating
Criteria
Adequate
• Evidence that exposure was consistently assessed using well-established methods
that directly measure exposure (e.g., measurement of PFAS in blood, serum, or
plasma), but there were some minor concerns about quality control measures or
other potential for nondifferential misclassification.
OR
• Exposure was assessed using indirect measures (e.g., drinking water concentrations
and residential location/history, questionnaire, or occupational exposure assessment
by a certified industrial hygienist) that have been validated or empirically shown to
be consistent with methods that directly measure exposure (i.e., intermethods
validation: one method vs. another). Note: This could be good if the validation was
sufficient. All studies for PFBS used direct measures.
And all of the following:
• Exposure was assessed in a relevant time window for development of the outcome.
• There is evidence that a sufficient proportion of the exposure data measurements are
above the limit of quantification for the assay.
• There is sufficient specificity/sensitivity and range or variation in exposure
measurements that would minimize potential for exposure measurement error and
misclassification by allowing exposure classifications to be differentiated (i.e., can
reliably categorize participants into groups such as high vs. low exposure), but there
might be more uncertainty than in good.
Deficient
Any of the following:
• Some concern, but no direct evidence, that the exposure was assessed using poorly
validated methods.
• There is insufficient information provided about the exposure assessment, including
precision, accuracy, and level of quantification, but no evidence for concern about
the method used.
• Exposure was assessed in a relevant time window for development of the outcome.
There could be concerns about reverse causation between exposure and outcome,
but there is no direct evidence that it is present.
• There is some concern over insufficient specificity /sensitivity and range or variation
in exposure measurements that may result in considerable exposure measurement
error and misclassification when exposure classifications are compared (i.e., data do
not lend themselves to reliably categorize participants into groups such as high vs.
low exposure, and/or there is considerable uncertainty in exposure values that do not
allow for confidence in the examination of small per unit changes in continuous
exposures).
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Table C-2. Criteria for Evaluation of Exposure Measurement in Epidemiology Studies
Exposure
Measurement Rating
Criteria
Critically deficient
Any of the following:
•	Exposure was assessed in a time window that is unknown or not relevant for
development of the outcome. This could be due to clear evidence of reverse
causation between exposure and outcome, or other concerns such as the lack of
temporal ordering of exposure and disease onset, insufficient latency, or having
exposure measurements that are not reliable measures of exposure during the
etiologic window.
•	Direct evidence that bias was likely because the exposure was assessed using
methods with poor validity.
•	Evidence of differential exposure misclassification (e.g., differential recall of
self-reported exposure).
•	There is evidence that an insufficient proportion of the exposure data measurements
are above the limit of quantification for the assay.
PFAS = per- and polyfluoroalkyl substances; PFBS = perfluorobutane sulfonic acid.
100

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Table C-3. Questions Used to Guide the Development of Criteria for Each Domain in Experimental Animal Toxicology Studies
Evaluation
Type
Domain-
Core Question
Prompting Questions
Basic Considerations
Reporting Quality
Reporting quality-
Does the study report
information for evaluating the
design and conduct of the study
for the endpoint(s)/outcome(s)
of interest?
Notes:
Reviewers should reach out to
study authors to obtain missing
information when studies are
considered key for hazard
evaluation and/or
dose-response.
This domain is limited to
reporting. Other aspects of the
exposure methods, experimental
design, and endpoint evaluation
methods are evaluated using the
domains related to risk of bias
and study sensitivity.
Does the study report the following?
• Critical information necessary to perform
study evaluation:
o Species, test article name, levels and
duration of exposure, route (e.g., oral,
inhalation), qualitative or quantitative
results for at least one endpoint of interest.
• Important information for evaluating the
study methods:
o Test animal: strain, sex, source, and
general husbandry procedures.
o Exposure methods: source, purity, method
of administration.
o Experimental design: frequency of
exposure, animal age, and life stage during
exposure and at endpoint/outcome
evaluation.
o Endpoint evaluation methods: assays or
procedures used to measure the
endpoints/outcomes of interest.
These considerations typically do not need to be refined by
assessment teams, although in some instances the imDortant
information mav be refined depending on the
endpoints/outcomes of interest or the chemical under
investigation.
A judgment and rationale for this domain should be given for the
study. Typically, these will not change regardless of the
endpoints/outcomes investigated by the study. In the rationale,
reviewers should indicate whether the study adhered to GLP,
OECD, or other testing guidelines.
• Good: All critical and important information is reported
or inferable for the endpoints/outcomes of interest.
• Adeem ate: All critical information is reported but some
important information is missing. However, the missing
information is not expected to significantly impact the
study evaluation.
• Deficient:. All critical information is reported but
important information is missing that is expected to
significantly reduce the ability to evaluate the study.
• Critically deficient:. Study report is missing any pieces of
critical information. Studies that are critically deficient
for reporting are uninfonnative for the overall rating and
considered no further for evidence synthesis and
integration.
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Table C-3. Questions Used to Guide the Development of Criteria for Each Domain in Experimental Animal Toxicology Studies
Evaluation
Domain-

Type
Core Question
Prompting Questions
Basic Considerations

Allocation-
For each study:
These considerations typically do not need to be refined by

Were animals assigned to
experimental groups using a
method that minimizes selection
bias?
• Did each animal or litter have an equal
chance of being assigned to any
experimental group (i.e., random
allocation)?
assessment teams.

A judgment and rationale for this domain should be given for
each cohort or experiment in the study.
• Good. Experimental groups were randomized, and any

.2
• Is the allocation method described?
specific randomization procedure was described or

3
4)

• Aside from randomization, were any steps
inferable (e.g., computer-generated scheme). (Note that

SJ
g

taken to balance variables across
normalization is not the same as randomization [see
%

experimental groups during allocation?
response for adequate].)
S3
s
1

randomized but do not describe the specific procedure used
p*
9*1
Sm

(e.g., "animals were randomized"). Alternatively, the
5
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e
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u
—
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CO

study authors used a nonrandom method to control for
important modifying factors across experimental groups
(e.g., body-weight normalization).
• Not reported (interpreted as deficient): No indication of
randomization of groups or other methods
(e.g., normalization) to control for important modifying
factors across experimental groups.
• Critically deficient:. Bias in the animal allocations was
reported or inferable.
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Table C-3. Questions Used to Guide the Development of Criteria for Each Domain in Experimental Animal Toxicology Studies
Evaluation
Type
Domain-
Core Question
Prompting Questions
Basic Considerations
CS
s
O
ga
2
Observational bias/blinding-
Did the study implement
measures to reduce
observational bias?
.2
2
a
u
QJ
A
-a
s
e
o
o
CO
For each endpoint/outcome or grouping of
endpoints/outcomes in a study:
• Does the study report blinding or other
methods/procedures for reducing
observational bias?
• If not, did the study use a design or
approach for which such procedures can be
inferred?
• What is the expected impact of failure to
implement (or report implementation) of
these methods/procedures on results?
These considerations typically do not need to be refined by the
assessment teams. (Note that it can be useful for teams to identify
highly subjective measures of endpoints/outcomes where
observational bias may strongly influence results prior to
performing evaluations.)
A judgment and rationale for this domain should be given for
each endpoint/outcome or group of endpoints/outcomes
investigated in the study.
• Good. Measures to reduce observational bias were
described (e.g., blinding to conceal treatment groups during
endpoint evaluation; consensus-based evaluations of
histopathology lesions).3
• Adequate: Methods for reducing observational bias
(e.g., blinding) can be inferred or were reported but
described incompletely.
• Not reported. Measures to reduce observational bias were
not described.
o Interpreted as adequate—The potential concern for bias
was mitigated based on use of automated/computer-driven
systems; standard laboratory kits; relatively simple,
objective measures (e.g., body or tissue weight); or
screening-level evaluations of histopathology.
o Interpreted as deficient—The potential impact on the
results is major (e.g., outcome measures are highly
subjective).
• Critically deficient:. Strong evidence for observational bias
that could have impacted results.
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Table C-3. Questions Used to Guide the Development of Criteria for Each Domain in Experimental Animal Toxicology Studies
Evaluation
Type
Domain-
Core Question
Prompting Questions
Basic Considerations
Risk of Bias
Confounding/
variable control
Confounding-
Are variables with the potential
to confound or modify results
controlled for and consistent
across all experimental groups?
For each study:
• Are there differences across the treatment
groups (e.g., co-exposures, vehicle, diet,
palatability, husbandry, health status, and
so forth) that could bias the results?
• If differences are identified, to what extent
are they expected to impact the results?
These considerations may need to be refined by assessment
teams, as the specific variables of concern can vary by experiment
or chemical.
A judgment and rationale for this domain should be given for
each cohort or experiment in the study, noting when the potential
for confounding is restricted to specific endpoints/outcomes.
• Good. Outside of the exposure of interest, variables that are
likely to confound or modify results appear to be controlled
for and consistent across experimental groups.
• Adequate: Some concern that variables that were likely to
confound or modify results were uncontrolled or
inconsistent across groups but are expected to have a
minimal impact on the results.
• Deficient:. Notable concern that potentially confounding
variables were uncontrolled or inconsistent across groups
and are expected to substantially impact the results.
• Critically deficient:. Confounding variables were presumed
to be uncontrolled or inconsistent across groups and are
expected to be a primary driver of the results.
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Table C-3. Questions Used to Guide the Development of Criteria for Each Domain in Experimental Animal Toxicology Studies
Evaluation
Type
Domain-
Core Question
Prompting Questions
Basic Considerations
Risk of Bias
Reporting and attrition bias
Selective reporting and
attrition-
Did the study report results for
all prespecified outcomes and
tested animals?
Note:
This domain does not consider
the appropriateness of the
analysis/results presentation.
This aspect of study quality is
evaluated in another domain.
For each study:
Selective reporting bias:
• Are all results presented for
endpoints/outcomes described in the
methods (see note)?
Attrition bias:
• Are all animals accounted for in the
results?
• If there are discrepancies, do study authors
provide an explanation (e.g., death or
unscheduled sacrifice during the study)?
• If unexplained results, omissions, and/or
attrition are identified, what is the expected
impact on the interpretation of the results?
These considerations typically do not need to be refined by
assessment teams.
A judgment and rationale for this domain should be given for
each cohort or experiment in the study.
• Good: Quantitative or qualitative results were reported for
all prespecified outcomes (explicitly stated or inferred),
exposure groups, and evaluation time points. Data not
reported in the primary article is available from
supplemental material. If results, omissions, or animal
attrition are identified, the study authors provide an
explanation, and these factors are not expected to impact
the interpretation of the results.
• Adequate: Quantitative or qualitative results are reported
for most prespecified outcomes (explicitly stated or
inferred), exposure groups, and evaluation time points.
Omissions and/or attrition are not explained but are not
expected to significantly impact the interpretation of the
results.
• Deficient:. Quantitative or qualitative results are missing for
many prespecified outcomes (explicitly stated or inferred),
exposure groups and evaluation time points and/or high
animal attrition; omissions and/or attrition are not
explained and may significantly impact the interpretation
of the results.
• Critically deficient:. Extensive results omission and/or
animal attrition is identified and prevents comparisons of
results across treatment groups.
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Table C-3. Questions Used to Guide the Development of Criteria for Each Domain in Experimental Animal Toxicology Studies
Evaluation
Type
Domain-
Core Question
Prompting Questions
Basic Considerations
Sensitivity
Exposure methods sensitivity
Chemical administration and
characterization-
Did the study adequately
characterize exposure to the
chemical of interest and the
exposure administration
methods?
Note:
Consideration of the
appropriateness of the route of
exposure is not evaluated at the
individual study level.
Relevance and utility of the
routes of exposure are
considered in the PECO criteria
for study inclusion and during
evidence synthesis.
For each study:
•	Does the study report the source, purity,
and/or composition (e.g., identity and
percent distribution of different isomers) of
the chemical? If not, can the purity and/or
composition be obtained from the supplier
(e.g., as reported on the website)?
•	Was independent analytical verification of
the test article purity and composition
performed?
•	Did the study authors take steps to ensure
the reported exposure levels were accurate?
o For inhalation studies: Were target
concentrations confirmed using reliable
analytical measurements in chamber air?
o For oral studies: If necessary, based on
consideration of chemical-specific
knowledge (e.g., instability in solution;
volatility) and/or exposure design (e.g., the
frequency and duration of exposure), were
chemical concentrations in the dosing
solutions or diet analytically confirmed?
•	Are there concerns about the methods used
to administer the chemical (e.g., inhalation
chamber type, gavage volume, etc.)?
It is essential that these criteria are considered and potentially
refined by assessment teams, as the specific variables of concern
can vary by chemical.
A judgment and rationale for this domain should be given for
each cohort or experiment in the study.
•	Good: Chemical administration and characterization is
complete (i.e., source, purity, and analytical verification of
the test article are provided). There are no concerns about
the composition, stability, or purity of the administered
chemical or the specific methods of administration. For
inhalation studies, chemical concentrations in the exposure
chambers are verified using reliable analytical methods.
•	Adequate: Some uncertainties in the chemical
administration and characterization are identified but these
are expected to have minimal impact on interpretation of
the results (e.g., source and vendor-reported purity are
presented, but not independently verified; purity of the test
article is suboptimal but not concerning). For inhalation
studies, actual exposure concentrations are missing or
verified with less reliable methods.
•	Deficient:. Uncertainties in the exposure characterization
are identified and expected to substantially impact the
results (e.g., source of the test article is not reported, levels
of impurities are substantial or concerning, deficient
administration methods such as use of static inhalation
chambers or a gavage volume considered too large for the
species and/or life stage at exposure).
•	Critically deficient:. Uncertainties in the exposure
characterization are identified, and there is reasonable
certainty that the results are largely attributable to factors
other than exposure to the chemical of interest
(e.g., identified impurities are expected to be a primary
driver of the results).
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Table C-3. Questions Used to Guide the Development of Criteria for Each Domain in Experimental Animal Toxicology Studies
Evaluation
Type
Domain-
Core Question
Prompting Questions
Basic Considerations
Sensitivity
Exposure methods sensitivity
Exposure timing, frequency
and duration-
Was the timing, frequency, and
duration of exposure sensitive
for the endpoint(s)/outcome(s)
of interest?
For each endpoint/outcome or grouping of
endpoints/outcomes in a study:
• Does the exposure period include the
critical window of sensitivity?
• Was the duration and frequency of
exposure sensitive for detecting the
endpoint of interest?
Considerations for this domain are highly variable depending on
the endpoint(s)/outcome(s) of interest and must be refined by
assessment teams.
A judgment and rationale for this domain should be given for
each endpoint/outcome or group of endpoints/outcomes
investigated in the study.
• Good. The duration and frequency of the exposure was
sensitive, and the exposure included the critical window of
sensitivity (if known).
• Adequate: The duration and frequency of the exposure was
sensitive, and the exposure covered most of the critical
window of sensitivity (if known).
• Deficient:. The duration and/or frequency of the exposure is
not sensitive and did not include the majority of the critical
window of sensitivity (if known). These limitations are
expected to bias the results towards the null.
• Critically Deficient: The exposure design was not sensitive
and is expected to strongly bias the results towards the null.
The rationale should indicate the specific concern(s).
107

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Table C-3. Questions Used to Guide the Development of Criteria for Each Domain in Experimental Animal Toxicology Studies
Evaluation
Type
Domain-
Core Question
Prompting Questions
Basic Considerations
Sensitivity
Outcome measures and results display
Endpoint sensitivity and
specificity-
Are the procedures sensitive and
specific for evaluating the
endpoint(s)/outcome(s) of
interest?
Note:
Sample size alone is not a
reason to conclude an
individual study is critically
deficient.
For each endpoint/outcome or grouping of
endpoints/outcomes in a study:
•	Are there concerns regarding the specificity
and validity of the protocols?
•	Are there serious concerns regarding the
sample size (see note)?
•	Are there concerns regarding the timing of
the endpoint assessment?
Considerations for this domain are highly variable depending on
the endpoint(s)/outcome(s) of interest and must be refined by
assessment teams.
A judgment and rationale for this domain should be given for
each endpoint/outcome or group of endpoints/outcomes
investigated in the study.
Examples of potential concerns include:
•	Selection of protocols that are insensitive or nonspecific for
the endpoint of interest.
•	Use of unreliable methods to assess the outcome.
•	Assessment of endpoints at inappropriate or insensitive
ages, or without addressing known endpoint variation
(e.g., due to circadian rhythms, estrous cyclicity, etc.).
•	Decreased specificity or sensitivity of the response due to
the timing of endpoint evaluation, as compared to exposure
(e.g., short-acting depressant or irritant effects of
chemicals, insensitivity due to prolonged period of
nonexposure prior to testing).
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Table C-3. Questions Used to Guide the Development of Criteria for Each Domain in Experimental Animal Toxicology Studies
Evaluation
Type
Domain-
Core Question
Prompting Questions
Basic Considerations
Sensitivity
Outcome measures and results display
Results presentation-
Are the results presented in a
way that makes the data usable
and transparent?
For each endpoint/outcome or grouping of
endpoints/outcomes in a study:
•	Does the level of detail allow for an
informed interpretation of the results?
•	Are the data analyzed, compared, or
presented in a way that is inappropriate or
misleading?
Considerations for this domain are highly variable depending on
the outcomes of interest and must be refined by assessment teams.
A judgment and rationale for this domain should be given for
each endpoint/outcome or group of endpoints/outcomes
investigated in the study.
Examples of potential concerns include:
•	Nonpreferred presentation such as developmental toxicity
data averaged across pups in a treatment group when litter
responses are more appropriate.
•	Failure to present quantitative results.
•	Pooled data when responses are known or expected to
differ substantially (e.g., across sexes or ages).
•	Failure to report on or address overt toxicity when
exposure levels are known or expected to be highly toxic.
•	Lack of full presentation of the data (e.g., presentation of
mean without variance data; concurrent control data are not
presented).
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Table C-3. Questions Used to Guide the Development of Criteria for Each Domain in Experimental Animal Toxicology Studies
Evaluation
Type
Domain-
Core Question
Prompting Questions
Basic Considerations
i
J
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Overall Confidence-
Considering the identified
strengths and limitations, what
is the overall confidence rating
for the endpoint(s)/outcome(s)
of interest?
Note:
Reviewers should mark studies
that are rated lower than high
confidence only due to low
sensitivity (i.e., bias towards the
null) for additional
consideration during evidence
synthesis. If the study is
otherwise well-conducted and
an effect is observed, the
confidence may be increased.
For each endpoint/outcome or grouping of
endpoints/outcomes in a study:
•	Were concerns (i.e., limitations or
uncertainties) related to the reporting
quality, risk of bias, or sensitivity
identified?
•	If yes, what is their expected impact on the
overall interpretation of the reliability and
validity of the study results, including
(when possible) interpretations of impacts
on the magnitude or direction of the
reported effects?
The overall confidence rating considers the likely impact of the
noted concerns (i.e., limitations or uncertainties) in reporting,
bias, and sensitivity on the results.
A confidence rating and rationale should be given for each
endpoint/outcome or group of endpoints/outcomes investigated in
the study.
•	High confidence: No notable concerns are identified
(e.g., most or all domains rated good).
•	Medium confidence: Some concerns are identified, but
expected to have minimal impact on the interpretation of
the results (e.g., most domains rated adequate or good;
may include studies with deficient ratings if concerns are
not expected to strongly impact the magnitude or direction
of the results). Any important concerns should be carried
forward to evidence synthesis.
•	Low confidence: Identified concerns are expected to
significantly impact the study results or their interpretation
(e.g., generally, deficient ratings for one or more domains).
The concerns leading to this confidence judgment must be
carried forward to evidence synthesis (see note).
•	Uninformative: Serious flaw(s) that make the study results
unusable for informing hazard identification
(e.g., generally, critically deficient rating in any domain;
many deficient ratings). Uninformative studies are
considered no further in the synthesis and integration of
evidence.
Tor nontargeted or screening-level histopathology outcomes often used in guideline studies, blinding during the initial evaluation of tissues is generally not
recommended because masked evaluation can make "the task of separating treatment-related changes from normal variation more difficult" and "there is concern
that masked review during the initial evaluation may result in missing subtle lesions." Generally, blinded evaluations are recommended for targeted secondary
review of specific tissues or in instances when there is a predefined set of outcomes that is known or predicted to occur (Crissman et at. 20041.
GLP = Good Laboratory Practice; OECD = Organisation for Economic Cooperation and Development.
110

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APPENDIX D. HAWC USER GUIDE AND FREQUENTLY ASKED QUESTIONS
D.l WHAT IS HAWC AND WHAT IS ITS PURPOSE?
The Health Assessment Workspace Collaborative (HAWC) is an interactive,
expert-driven, content management system for human health assessments that is intended to
promote transparency, trackability, data usability, and understanding of the data and decisions
supporting an environmental and human health assessment. Specifically, HAWC is an interface
that allows the data and decisions supporting an assessment to be managed in modules
(e.g., study evaluation, summary study data, etc.) that can be publicly accessed online (see
Section D.2 below and Figure D-l). Following the literature search and screening that are
conducted using HERO and DistillerSR. HAWC manages each study included in an assessment
and makes the extracted information available via a web link that takes a user to a web page
displaying study-specific details and data (e.g., study evaluation, experimental design, dosing
regime, endpoints evaluated, dose-response data, etc., described in further detail below in
Sections D.3 to D.6). Finally, all data managed in HAWC is fully downloadable using the blue
"Download datasets" link (highlighted in the red box below) also located in the gray navigation
bar located on the assessment home page (discussed in Section D.7). Note that a user may
quickly navigate HAWC by clicking on the file path (highlighted in the orange, dashed box
below) given in the gray row below the HAWC icon and menu bar (see Figure D-l). HAWC
aims to facilitate team collaboration by scientists who develop these assessments and enhance
transparency of the process by providing online access (no user account required) to the data and
expert decisions used to evaluate potential human health hazard and risk of chemical exposures.
- » ~ e :
Figure D-l. HAWC Homepage for the Public PFBS Assessment
D.2 HOW DO I ACCESS HAWC?
HAWC is an open-source, online application that may be accessed using the following
link—https://hawcprd.epa.gov/assessment/public/—and then selecting an available assessment.
The following browsers are fully supported for accessing HAWC: Google Chrome (preferred),
Mozilla Firefox, and Apple Safari. There are errors in functionality when viewed with Internet
Explorer. No user account is required for access to public HAWC assessments. The
assessments located in HAWC are meant to accompany a textual expert synthesis of the data
managed in HAWC. Each written assessment document contains embedded URL links to the
evidence in HAWC (e.g., study evaluation, summary study data, visualizations, etc.) supporting
<- C O i Secure https hawcprd.epa.gov assessment/100000037/
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111

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EPA 690 R-21 00 IP
the assessment text. The links embedded in an assessment document can be accessed by a
mouse click (or hover while pressing CTRL + right click).
D.3 WHAT CAN I FIND IN HAWC?
HAWC contains a comprehensive landscape of study details and data supporting an
assessment. Note that links are provided in the assessment text to guide the reader, but a user
may also navigate to the HAWC homepage for an assessment 011 their own. Once a user lands
on an assessment homepage, all studies included in an assessment can be viewed by clicking the
blue "Study list" link (highlighted in the red box below) in the gray navigation pane (see Figure
D-2). By clicking the study name listed in blue (under "Short citation") a user can view the full
study details, study evaluation, and experimental details and data. For example, in Figure D-2, a
user may click on "3M, 2000, 4289992" (highlighted in the orange, dashed box below). This
will take the user to the 3M (2000d) study details page that includes a link to the study in HERO
along with study details, study evaluation, and available experimental (animal) and study
population (epidemiologic) groups.
<- -> C O i Secure httpsV hawcprd.epa.gov st jdy/assessment/100000037/
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Public Assessments PFBS (2018) Stu
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Contact About Pubic Assessments logm
SELiCttO ASSESSMENT
Available studies
Animal btoassay
Epidenvoiogy
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Short r.itatKwi
3M. 2000. 4289992
3M. 2001.4241248
3M. 2010.3927382
Bao, 2017.3860099
Berk, 2014,2713574
B.y*nd, 2011.1578502
Bombard. 1996.3859928
Full citation
A repeated dose range-finding towiy study of T-748S in Sprague-Dawfey ret#. ~
STUDY NUMBER: 132-006. SPONSOR: 3M Pharmaceutical, St Paul, MN
55133-3320. TESTING FACILITY Pnmedlca Redfiotd, Redfieid, AR 72132 STUDY
DATES Study Initiation: Juno 26.2000 Animal Phase Initiation: June 27,2000
Animal Phase CompSetion: July 7, 2000 Study Completion: October 11,2000
3M. A 28-day oral (gavage) toxicity study of T-7485 In Sprague-Oawtey rats. V
[Study Number 132-007). St. Paufl. MN: 3M Corporate Toxicology.
3M. TSCA 8(e) Substanlial Risk Note®: Sulfonate-based and Carboxylic-basod v
Fluorochem
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EPA 690 R-21 00IF
rating. A user may hover over each piece of the pie, which causes rating metric text to populate
to the right of the pie graph (see Figure D-3). For full domain and rating details the user may
click the blue "View details" button (highlighted in the red box below). [Note that this example
is given for the 3M (2000d)1.
Risk of bias visualization
Figure D-3. Representative Study Evaluation Pie Chart with the Reporting Domain
Selected and Text Populating to the Right of Pie Chart
D.5 HOW DO I ACCESS STUDY-SPECIFIC INFORMATION ON EXPERIMENTAL
AND STUDY POPULATION DETAILS AND EXTRACTED ENDPOINT DATA?
Specific information on experimental design, dosing (if animal bioassay), outcomes and
exposure (if epidemiology), and extracted endpoint data can be accessed from the study details
page by clicking on [for the 3M (2000d) study] "available animal bioassay experiments" at the
bottom of the study details page. A user may click on the experiment name (highlighted in blue,
10-dav oral) to view dosing/exposure details and available groups. Clicking on available animal
groups (e.g., male Sprague-Dawley or female Sprague-Dawley) will take the reader to a new
page with experimental group information (e.g., species/strain/sex, dosing regimen information,
and available/additional endpoints information for animal studies; and outcome and exposure
information for epidemiologic studies). If a study reports data, then the data are extracted and
managed as "available endpoints." If the study authors include endpoints in the methods and
results but do not report data, the endpoint is listed under "additional endpoints" without
dose-response data. All endpoints are also clickable and contain an endpoint description,
methods, and (if data are reported) a clickable data plot (e.g., alanine aminotransferase IALT"!).
The description of endpoints, methods, and data are often copied directly from the study report
and, therefore, can contain study author judgments and may not necessarily include U.S. EPA
judgments on the endpoint data that would be included in the assessment.
Reporting domain
^9 Reporting of information necessary for study
evaluation
Good. Important information is provided for test
species, strain, source, sex, exposure methods,
experimental design, endpoint evaluations and the
presentation of results. The study was not conducted
under GLP guidelines; however, several reviews were
performed by the Quality Assurance Unit.
3M, 2000, 4289992 risk of bias summary
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D.6 WHAT ARE VISUALIZATIONS AND HOW DO I ACCESS THEM?
The data managed in HAWC is displayed using visualizations that are intended to
support textual descriptions within an assessment. All visualizations can be accessed using the
blue "Visualizations" link (highlighted in the red box below) also found in the gray navigation
pane (see Figure D-4A). Note that the available visualizations are at the discretion of the
chemical manager and are meant to accompany the assessment text. Visualizations are fully
interactive. Hovering and clicking on records in the rows and columns and data points on a plot
will cause a pop-up window to appear (see Figure D-5B). This pop-up window is also
interactive and clicking on blue text within this pop-up will open a new web page with
descriptive data.
Assessments PFBS <2Qta> Visualizations PFBS T4 (affect size, snir
Contact About Public Assessments Logan
PFBS T4 (effect size, animal)
AVAILABLE MODULES
gov. w 0*55
Download datawte
Figure D-4A. Visualization Example for PFBS
(Note that the records listed under each column [study, experiment endpoint, units, study
design, observation time, dose] and data within the plot are interactive.)
114
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EPA 690 R-21 00 IP
NTP 2018, 4309741 / 28 Day Oral / Male Harlan Sprague Dawley Rat / Tetraiodothyronine (T4), Free
Study Experiment Animal Group Endpoint
Endpoint name
Tetraiodothyronine (T4), Free
System
Endocrine
Organ
Thyroid
Effect
Hormone
Effect subtype
Thyroid Hormone
Observation time
Day 28
Data reported?
V
Data extracted?
<~
Values estimated?
-
Location in literature
R07-Hormone Summary
Expected response
adversity direction
any change from reference/control group
LOAEL
62.6 mg/kg-day
Monotonicity
not-reported
Statistical test description
Jonchkeere (trend) and Shirley or Dunn (pairwise) tests
Trend result
not reported
Power notes
Statistical analysis performed by Jonchkeere (trend) and Shirley or Dunn (pairwise) tests Statistical significance for a
treatment group indicates a significant pairwise test compared to the vehicle control group Statistical significance for the
control group indicates a significant trend test * Statistically significant at P <= 0.05 '* Statistically significant at P <= 0.01
Figure D-4B. Example Pop-Up Window after Clicking on Interactive Visualization Links
(In Figure D-4A, the red circle for study NTP (2019): male at a dose of 500 ing/kg-day was
clicked leading to the pop-up shown above. Clicking on the blue text will open a new
window with descriptive data.)
D.7 HOW DO I DOWNLOAD DATA SETS?
A user may download any available data set by first clicking on the blue "Download
data sets" link (highlighted in the red box below) in the gray navigation pane on the assessment
homepage. This takes the user to a new page where the desired data set may be selected for
download as an Excel file (see representative image in Figure D-6).
115
-------
EPA 690 R-21/00 IF
f 0 Cs i Secure https /hawcprd.epa.gov assessment/l 00000037/downloads/
Q ^	Upgrade now
space and she
Contact About Public Assessments Login
Public Assessments PFBS{201B) Downloads
SELECTED ASSESSMENT
AVAILABLE MODULES
Study list
Visualizations
Download dalaseis
PFBS (2018) downloads
ah data !rom hawc *"» exportable into E«ei. Devofoper exports m JSON 
-------
EPA 690 R-21 00 IP
0 i Secure- https ¦¦ hawcprd.epa.gov /ani/endpoint..'! aOGO,?J8S/
Contact About Public Assessments Login
PuWic Assessments PFBS (2018) Feng, 2017.3856465 20 Day Oral Gestation P0 Femafe ICR Mice Teira#odoiby!c
SELECTED ASSESSMENT
AVAILABLE MODULES
Study list
Visualizations
DOWNLOADS
Download ciaiasels
Tetraiodothyronine (T4), Free
Endpoint Details
Endpoint name
Tetraiodothyronine (T4), Free
System
Endocrine
Organ
Thyroid
Effect
Hormone
Effect subty pe
Thyroid Hormone
Observation time
Additional tags
GD20
Wgti confidence
Data reported?
~
Data extracted?
~
Values estimated? —
Location in literature
Table 3
NOAEL
50 mg/Rg-day
Plot
BMP Modeling
Tctralodotfiyfonii
View session
U!he- opfiins
Qo&ses *i Study
^LOAEL
^NOA£l.
{
too 200 300 -too 500
Dose (mftfcg-day)	
Figure D-6A. Example BMD Modeling Navigation
C i Secure hups ¦' Tiaw«prd,epa.gov. :ii--in'/se',s.ion/iOOOOC Number of Animals
Response {pg/ml)
Standard Error
0 8
16.81
0.7
7.8® a
17.58
0.64
31*° 8
14.74
0.51
77b 8
14.95
0.46
" NOAEL (No ObseiYW Ahorse Eltecl
" S^jnttainOy different from control (p < 0.05)
c LOAEL 4Low&st Observed Adverse Efiecs Leve
Selected models and options
0MDS V2.7.0
Model options
Model nam® Non-default settings
Tetralotfothyronlne (T4|, Free

m study
#LOA£L
•nOAEL
I i
Dose im§> tig-day HEO)
i.
Benchmark modeling responses
Confidence
Figure D-6B. Example BMD Session
117
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EPA 690 R-21 00 IP
APPENDIX E. ADDITIONAL DATA FIGURES
Experiment Endpoint
l_ nits Sludy Design
Observation Confidence
Dose
(mg/kg-day)
NTP2018	28-day oral Free Telraiodothyroninc 
-------
EPA 690 R-21 00 IP
Route	l'\|xiMirc K ml point
Confidence	Units
Feng, 2017. 3856465 H Mouse. ICR ($, N=10)
oral gavage GDI to 20 Triiodothyronine (T3)
Dost
(mg/kg-day)
NTP, 2018.4309741 Rat, Harlan Sprague-Dawley ($. N=l-10) oral gavage 28 days Triiodothyronine (T3)	Ihigh confidencel ng/dl. 0
Rat. Harlan Sprague-Dawley (Cf. N=9-10) oral gavage 28 days Triiodothyronine ("I'll)
Fl Mouse. ICR (9- N=30)
oral gavage
GDI t
o 20
Triiodothyronine (T3)
ng/ml
0






50






200
500
P0 Mouse. ICR (9, N=8)
oral gavage
GDI t
o 20
Triiodothyronine (T3)
ng/ml
0
50
200
500
Fl Mouse. ICR (9. N=I0)
oral gavage
GDI t
o 20
Triiodothyronine (T3) - Litter N
ng/ml
0
-60 -40 -20 0 20
percent control response
Figure E-2. Serum Total Triiodothyronine (T.<) Response in Animals Following
K+PFBS Exposure
(Click to see interactive data graphic)
119
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EPA 690 R-21 00 IP
Study
Study Design
Exposure
observation time
Units
Dose
(mg/kg-day)
NTP 2018.4309741
Rat. Harlan Spraguc-Dawlcy (9. N=l-10)
28 days
Day 28
ng/mL
0
62.6
125
250





500





1,000

Rat. Harlan Spraguc-Dawley (cf, N=9-10)
28 days
Day 28
ng/mL
0
62.6
125





250





500
Feng 2017. 3856465
F1 Mouse, ICR ($, N=10)
GDI to 20
PND1
ng/ml
0
50
200
500



PND30
ng/ml
0
50
200





500



PND60
ng/ml
0





50





200
500

P0 Mouse, ICR (9, N=8)
GDI to20
C.D20
ng/ml
0
D statistically significant
£ percent control response
IH 95% CI
200
500













<






















H
1
1—0—1

H
!-•




















-100 -80 -60 -40 -20 0 20 40 60 80 100
percent control response
Figure E-3. Serum Thyroid-Stimulating Hormone (TSII) Response in Animals Following
K+PFBS Exposure
(Click to see interactive data graphic)
Study	Experiment Endpoint	Units Study Design	observation Confidence Dose
time	(mg.'kg-day)
Feng. 2017, 3856465 2CMjayorai Eye Opening
days F1 Mouse, ICR (2. N=5Q) Beginning o
PND12
Eye Opening - Litter N days F1 Mouse, CR (£. N=10) Beginning on
PND12
50
200
500
0
50
200
5CO
El statistically significant
0 percent control: resocrs
MS5% CI
	1	
*
*
' «
i
i h
•
i
-60-40-20 0 20 40 00
peroenl contra' response
Figure E-4. Developmental Effects (Eye Opening) Following K+PFBS Exposure in Rats
(Click to see interactive data graphic)
120
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EPA 690 R-21 00 IP
Study
Study Design
Route	Exposure Kndpoint
Units Dose
(mg/kg-day)
O statistically signific
£ percent control resj
H 95% CI
Feng, 2017, 3856465 F1 Mouse. ICR (9. N=30) oral gavage GDI to 20 Estrous Cycle. Diestrus days 0
50
200
500
FI Mouse. ICR (9. N= 10) oral gavage GDllo20 Estrous Cycle, Proestrus days 0
50
200
500
M H
~I	1	1 I
©
~i	1	1	r~
-100 -80 -60 -40 -20 0 20 40 60 80 100
percent control response
Figure E-5. Developmental Effects (First Estrus) Following K+PFBS Exposure in Rats
(Click to see interactive data graphic)
Route	Exposure Kndpoint
Units Dose
(mg/kg-day)
Lieder. 2009, 1578545 Fl Rat, Sprague-Dawley (9, N=30) oral gavage	Vaginal Patency	days 0
30
300
1.000
Feng. 2017,3856465 Fl Mouse. ICR (9, N=30)	oral gavage GDI to 20 Vaginal Patency	days 0
200
500
Fl Mouse. ICR <9. N=I0)	oral gavage GDI to 20 Vaginal Patency (litter) days 0
200
500
O statistically significant
0 percent control response
M 95% CI
-60 -40 -20 0 20 40 60 80 100
percent control response
Figure E-6. Developmental Effects (Vaginal Patency) Following K+PFBS Exposure in Rats
(Click to see interactive data graphic)
121
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EPA 690 R-21 00 IP
Animal deMrriplian
Incidence	Dow
(Hiftflts-dayl
IcoittidefdJ
BmctdcfKr
Lietfer, 2009, 1578546 Rat, CrlrCd^lSdjIg* Br VatfPliKfcm (9. N»10) oral gavage Mibchroaie Kidney, Edema. Focal Pajnllary
(90
Kidney. HypetpUua, Papillary
TuMar/Diictal Epithelium
Rat, CrirCdWSdtfp Br Vaf/PluMm «f. N»I0)
Lieder. 2009.1578545 Fl Rai. Spraguc-Dawfcy 19. N=X»
al gavage Mibchronic Kidney, Edema. Focal Papillary
(90 day*)
Kidney, Hyperplasia. Papillary
Tubular,'Ductal Epithelium
oral gavoge rcjiroduciive Kidney. Edema. Focal Papillary
Fl Rai. Sprague-Dawfcv iff. N=30l
Kidney. Ilypeiptasta, Papillary
Tubular/Ductal Epithelium
oral gavage reproduclivi* Kidney, Edema. Focal Papillary
P0 Rut, Sproguc-Dawley (9- N»30)
Kidney, tlypi'ljiUvti, Papillary
Tul*ilar,it>iictal Epithelium
«al gavage reproductive Kidney, Edema. Focal Papillary
It) Rat. Sprague-Dawley 
-------
EPA 690 R-21 00 IP
l%iul point Name
Study Nunc
F.xperiment Name
Animal Description
Observation Time
Blimd Urea Nilrogen (BUN)
3M, 2000,4289992
10 Day Oral
Rat, CH: Cd (Sd) lbs Br (cf)
Day 11



Ral. Crl: Cd (Sd) lbs Br (9)
Day 11

NTP 2018, 4309741
28 Day Oral
Rat, Harlan Spraguc-Dawlcy (Cf)
Day 28



Rat, Harlan Sptaguc-Dawlcy (9)
Day 28

3M. 2001. 4241246
28 Day Oral
Rat, Cri:CD(SD) (Cf)
Day 29



Rat. CrtCDlSDI (9)
Day 29

Lieder. 2009. 1578546
90 Day Oral
Rat. Cri:Cd@(Sd)lgs Br Vaf/Plustm (cf)
Day 90



Rat. Crl:Cd@(Sd)lgs Br Vaf/Plustm (9>
Day 90
Creatinine (CREAT)
3M, 2000, 4289992
10 Day Oral
Rat, CH: Cd (Sd) lbs Br 1 cf)
Day 11



Ral. CH: Cd (Sd) lbs Br (9)
Day 11

NTP 2018. 4309741
28 Day Oral
Ral. Harlan Spraguc-Dawlcy {(f)
Day 28



Rat. Harlan Spraguc-Dawlcy 19)
Day 28

3M. 2001.4241246
28 Day Oral
Rat, Crl:CD(SD) (cf)
Day 29



Rat. Cri:CD(SD) (9)
Day 29

Lieder. 2009. 1578546
90 Day Oral
Rat. Crl:Cd@(Sd)lgs Br Vaf/Plustm ICf)
Day 90



Rat. Crl:Cd@(Sd)lgs Br Vaf/Plustm (9)
Day 90
Hydronephrosis
3M. 2001.4241246
28 Day Oral
Rat. CrlrCDfSD) (Cf)
Day 29




Day 29



Rat, Crt:CD(SD) (9)
Day 29
Kidney. Hydronephrosis
3M. 2000, 4289992
10 Day Oral
Rat. Crl: Cd (Sd) lbs Br (Cf)
Day 11



Rat. Crl: Cd (Sd) lbs Br (9)
Day 15
Kidney. Cortical Tubular Dilation. Focal
Lieder, 2009. 1578545
2 Generation Oral
P0 Rat. Spraguc-Dawlcy (Cf 1




P0 Rat. Sprague-Dawley (9)
LD22



Fl Rat. Sprague-Dawley (cf)
Day 120



FI Rat. Sprague-Dawley (9)
Day 120
Kidney. Right. Pelvis. Dilation
Lieder. 2009, 1578545
2 Generation Oral
Fl Rat. Sprague-Dawley (Cf9)
LD 22
Kidney. Right, Pelvis, Marked Dilation
Lieder. 2009. 1578545
2 Generation Oral
FI Rat. Spraguc-Dawlcy (Cf)

Kidney. Right. Pelvis. Slight Dilation
Lieder. 2009. 1578545
2 Generation Oral
PI Rat. Spraguc-Dawlcy (cf)

Kidney. Edema, Focal Papillary
Licdcr. 2009. 1578545
2 Generation Oral
P0 Rat, Sprague-Dawley (Cf)




PU Rat. Sprague-Dawley (9)
LD 22



FI Rat. Sprague-Dawley (Cf)
Day 120



Fl Rat. Sprague-Dawley (9)
Day 120

Lieder. 2009. 1578546
90 Day Oral
Rat, Crl:Cd(Sd)Igs Br Vaf/Plustm tcf)
Day 90



RaU Crl:Cd@(Sd)!gs Br Vaf/PlusUn (9)
Day 90
Kidney Basophilia. Tubular, Multifocal
Lieder, 2009, 1578546
90 Day Oral
Rat, Crl:Cd@(Sd)lgs Br Vaf/Plustm (Cf)
Day 90



Rat. Crl:Cd@(Sd)Igs Br Vaf/Plustm <9>
Day 90
Subacute Inflammation. Cortex
3M. 2001.4241246
28 Day Oral
Rat. CihCD(SD) (Cf)
Day 29
Kidney. Mononuclear Cell Infiltrate
Lieder, 2009, 1578546
90 Day Oral
Rat. Cri:Cd®(Sd)lgs Br Vaf/Plustm tcf)
Day 90



Rat Cd:Cd©(Sd)lgs Br Vaf/Plustm (9)
Day 90
Kidney. Pyelonephritis. Chronic
Lieder. 2009. 1578546
90 Day Oral
Rat. CrI:Cd®(Sd)lgs Br Vaf/Plustm (cf)
Day 90



Rau Crl:Cd@(Sd)lgs Br Vaf/Plusun (9)
Day 90
Kidney. Necrosis, Pupillary
Lieder. 2009, 1578545
2 Generation Oral
PO Rat, Sprague-Dawley (cf)




PO Rat. Sprague-Dawley (9)
LD 22



Fl Rat, Spraguc-Dawlcy (Cf 1
Day 120



FI Rat. Sprague-Dawley (9)
Day 120

NTP 2018,4309741
28 Day Oral
Rat. Harlan Sprague-Dawley (cf)
Day 28



Rat, Harlan Sprague-Dawley (9)
Day 28

Lieder. 2009. 1578546
90 Day Oral
Rat. CH:Cd@(Sd)lgs Br Vaf/Plusun (cf)
Day 90



Rau Crl:Cd@(Sd)Igs Br Vaf/Plustm (9)
Day 90
Kidney. Tubular Degeneration
3M, 2000, 4289992
10 Day Oral
Rat. CH: Cd (Sd) lbs Br tcf)
Day 11



Ral. Crl: Cd (Sd) lbs Br (9)
Day IS

3M. 2001. 4241246
28 Day Oral
Ral, CrirCD(SD) (cf)
Day 29
Tubular Degeneration
3M, 2001,4241246
28 Day Oral
Rat CrlrCD(SD) (9)
Day 29
Tubular Degeneration. Terminal
3M, 2001,4241246
28 Day Oral
Rat. Crl:CD(SD> (9)
Day 29
Kidney. Hyperplasia, Papillary Tubular.'Ductal Epithelium
Lieder, 2009. 1578545
2 Generation Oral
PO Rat. Sprague-Dawley (cf)




PORal. Sprague-Dawley (9)
LD 22



Fl Rat, Sprague-Dawley (cf)
Day 120



Fl Rat, Sprague-Dawley (9)
Day 120

lieder. 2009, 1578546
90 Day Oral
Ral, Crl:Cd®(Sd)Igs Br Vaf/Plustm (cf)
Day 90



Ral, Crl:Cd<8>(Sd)Tgs Br Vaf/Plustm (9)
Day 90
Kidney, Mineralization
3M, 2000, 4289992
10 Day Oral
Rat, CH: Cd (Sd) Ths Br (Cf)
Day II



Rat. Crl: Cd (Sd) lbs Br (9)
Day 15

Lieder. 2009. 1578546
90 Day Oral
Rat. Crlfd@(Sd)Igs Br Vaf/Plustm tcf)
Day 90



Rat. Crl:Cd@(Sd)lgs Br Vaf/Plustm (9)
Day 90
Mineralization
3M. 2001,4241246
28 Day Oral
Rat, Crl:CD(SD) (9)
Day 29
Mineralization, Terminal
3M, 2001,4241246
28 Day Oral
Rat, Crl:CD(SD) (9)
Day 29
Kidney, Nephropathy. Chronic Progressive
NTP 2018, 4309741
28 Day Oral
Ral. Harlan Sprague-Dawley (cf)
Day 28



Rat, Harlan Spraguc-Dawlcy (9)
Day 28
Kidney. Hyaline Droplets, Cortical Tubules
Licdcr. 2009. 1578546
90 Day Oral
Rat, Crl;Cd(*(Sd)lgs Br Vaf/Plustm (Cf)
Day 90



Ral, CH:Cd#(Sd)lgs Br Vaf/Plustm (9)
Day 90
Kidney Hislopulbolugy
3M. 2000,4289992
10 Day Oral
Ral, CH: Cd (Sd) lbs Br ICf)
Day 11



Ral, CH: Cd (Sd) lbs Br (9)
Day II

3M, 2001,4241246
28 Day Oral
Rat, CrlrCDfSD) (cf)
Day 29



Rat, CrtrCD(SD) (9)
Day 29
1'IBS Kidney Effects
Dow Range
Sianilicant Increase
.Significant Decrease
100 200 300 400 500 600 700 800 900
Dose (mg/kg/duy)
Figure E-8, Renal Effects Following K+PFBS Exposure in Rats
(Click to see interactive data graphic)
123
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EPA 690 R-21 00 IP
Kndpoint Name
Study Name
Kvperiment Name
Animal Description
Observation l ime
Kidney Weight. Absolute
Lieder. 2009. 1578545
2 Generation Oral
F2 Ral. Sprague-Dawley (Cf9)


3M. 2000.4289992
10 Day Oral
Rat. Crl: Cd 9, 1578545
2 Generation Oral
Pll Ral, Sprague-Dawley (cf >
P0 Ral. Sprague-Dawley (9)
LD22



Fl Rat. Sprague-Dawley (cf)
Day 120



Fl Rat. Sprague Dawley (9)
Day 120
Kidney Weight, Relative
Lieder, 2009, 1578545
2 Generation Oral
F2 Ral, Sprague-Dawley (cf 9)


3M, 2000.4289992
10 Day Oral
Ral. Crl: Cd (Sd) lbs Br (d1)
Day 11



Rat. Crl: Cd (Sd) lbs Br (9)
Day 11

3M, 2001,4241246
28 Day Oral
Rat, Crl:ClXSD) (cf)
Day 29



Rat. Crl:CD(SD) (9)
Day 29

Lieder, 2009,1578546
90 Day Oral
Ral, Cri:CdW(Sd)lgs Br VatVPIustm (Cf)
Ral. Crf:Cd@(Sd)Igs Br Var/Plustm (9)
Day 90
Day 90
Kidney Weight. Right. Relative
Lieder. 2009.1578545
2 Generation Oral
P0 Ral. Sprague-Dawley (cf)
P0 Rat. Sprague-Dawley (9)
FI Ral. Sprague-Dawley (cf)
LD 22
Day 120



Fl Rat. Sprague-Dawley (9)
Day f20

NTP 2018.4309741
28 Day Oral
Rat, Harlan Sprague-Dawley (Cf)
Day 28



Rat, Harlan Sprague-Dawley (9)
Day 28
PFBS Kidney Weight Effects
AA A
• Doses
M Dose Range
A Significant Increase
Significant Decrease
200 300 400 500 600
Dose (nig/kg/day)
700 800 900 1.000 1.100
Figure E-9. Kidney-Weight Effects Following K PFBS Exposure in Rats
(Click to see interactive data graphic)
124
-------
EPA 690 R-21 00 IP
tndpoint Nunc
ik AmimHrartslcrave (ALT)
NTP 2018. 4309741
1M, 2(11(1,4241246
Experiment Name
II) Day Oral
28 Day Oral
28 Day Oral
Licdcr. 2009. 1578546 90 Day Oral
Animal Description
Rat.CM:Cd(Sd) lb- Hit9)
Rat Crl: Cd (Sd) lbs Bf(CfI
Rat. HarUn Sprdguc-Dawlcv (9)
Rat. Harlan Sprague-Dawley «f)
Rat. Crl:CD
Day 11



Rat. Crl: Cd (Sd) Ibv Br (9)
Duy 15
Inflammation. Subacute
3M. 2001,4241246
28 Day Oral
Rat. Crl:CDlSD)<9l
Day 29



Rat CrfcCDtSDl (rf)
Day 29
.Subacute Inflammation
3M. 2001.4241246
28 Day Oral
Rat CrWDtSDl (9)
Day 29



Rat CitCMSD) (cfl
Day 29
Day 43
Subacute Inflammation. Teiminal
3M. 2001. 4241246
28 Day Oral
Rat Crl:CD(SDl (9)
Rat OM'DtSD) (cfl
Duy 29
Duy 2S
Inflammation. Chronic. Focal/Multirocal
Lieder. >009. 1578546
90 Day Oral
Rat. Crl.-Cdt?lSdilgs Br Val/Plusun (9)
Rat. CrirCdSHSdllgs Br Val/Pluvtm 1(f)
Day 90
Day 90
Hepatocellular Hypertrophy
Licdcr. 2009. 1578545
2 Generation Oral
P0 Rat. Sprague-Dawley 1(f)


NTP 2018. 4309741
28 Day Oral
Rut. Ilarlan Spnigue-Dawley (9)
Duy 28



Rat. Harlan Spruguc-Dawlcy (Cf)
Day 28

Licdcr. 2009. 1578545
2 Generation Oral
Fl Rat Sprague-Dawley (9)
PI Rat Sprague-Dawley (cf)
Day 120
Day 120



P0 Rat. Sprague-Dawley (9)
1.1)22
Liver. Necrosis
NTP20I8. 4309741
28 Day Oral
Rat. Harlan Sprague-Dawley <9)
Day 28



Rat Hartan Spruguc-Dawlcy tcf J
Day 28
Necnwli
3M. 2001,4241246
28 Day Oral
Rat Crl:CD»SD) (tf 1
Duy 29
Necrosis. Terminal
3M. 2001.4241246
28 Day Oral
Rat CrtCDiSD) (Cfl
Day 29
Bile Duct Qst
NTP 2018.4309741
28 Day Oral
Rai. Harlan Sprague-Dawley (9)
Day 28



Rat. Harlan Spruguc-Dawlcy «f J
Day 28
llepatudiaphragmatic Nodule
NTT" 2018, 4309741
28 Day Oral
Rat. Harlan Sprague-Dawley (9)
Rat. Harlan Sptnguc-Dawlcy «fi
Day 28
Day 28
Hemorrhage
NTP 2018. 4309741
28 Day Oral
Rat. Ilurlan Spruguc-Duwlcy (9)
Duy 28
Liver Hlatopathnlogy
Liver Weight, Absolute
Liver Weight. Absolute
Liver Weight Relative
3M. 2000.4289992
3M. 3001. 4241246
Licdcr. 20O9, 1578545
3M. 2000.4289992
NTP20I8, 4309741
3M-2001.4241246
Lieder. 2009. 1578546
Licdcr. 2009. 1578545
NTP 2018. 4309741
Licdcr. 2009. 157854S
10 Day Oral
28 Day Oral
2 Generation Oral
28 Day Oral
28 Day Oral
2 Generation Oral
28 Day Oral
2 Generation Oral
Rat. Harlan Sprague-Dawley (cfl
RaL Crl: Cd (Sd) 11*. Br (9k
Rai. Crl: Cd (Sd) lbs Br (cf I
Rai. CrliClXSDI (9)
Rat. CHrCDtSD) «f I
F2 Rat, Sprague-Dawley l
-------
EPA 690 R-21 00 IP
Endpolnt Name	Study Name Experiment Name	Animal Description	Observation Time
A|H>lipo|mjtc"in A1 (ApoAl)
Bijlund, 2011.1578502
Subchronic Oral
Mouse. Apoe"3-Leklen.Celp (cf)
Week 4
M
|—| Dose Range
Bile Acid Secretion
Bijbuid.2011, 1578502
Subchronic Oral
Mouse. Apoc*3-Leidcn.Cclp (Cf)
Week 4

• Doses
Chnbseml KsterTrunsfer Protein (OK
IT) Bijland, 2011. 1578502
Suhchmnic Oral
Mouse. Apoe"3-i.cklen.Cclp 
Day II

V Significant Decrease



Ral. Crl; Cd (Sd) lbs Br (cf)
Day 11
•—a	a	•


3M. 2001.4241246
28 Day Oral
Rat. CrtCD(SD) (Cf)
Day 29



Bijland, 2011,1578502
Subchronic Oral
Mouse. Apoc"3-Leiden.Cclp (Cf)
Week 1



l.foder, 2009, 1578546
90 Day Oral
Ral. C'rK'd W(SdHgs Br VaPI'lusIm (9)
Day 90





Rat. CrlCd®(Sd)lgs Br VaPPhuim («f)
Day 90
~—•	•	~

Cholesterol Estet (CE)
Bijland, 2011. IS78502
Subchronic Oral
Mouse. Apoe"3Leklai,Celp (0"l
Week 4


ChulesleiuL Free
Bijlund, 2011.1578502
Subcluunk- Oral
Mouse, Apoe"3-Leklen.Cclp 
-------
EPA/690/R-21/001F
APPENDIX F. BENCHMARK DOSE MODELING RESULTS
F.l. MODELING OF NONCANCER ENDPOINTS
As discussed in the body of the report under "Derivation of Oral Reference Doses," the
endpoints selected for benchmark dose (BMD) modeling were incidence of renal papillary
epithelial tubular/ductal hyperplasia in rats from Lieder et al. (2009a) and Lieder et al. (2009b);
thyroid hormones in pregnant mice and offspring at Postnatal Days (PNDs) 1,30, and 60 from
Feng et al. (2017) and adult rats from NTP (2019); and developmental effects (i.e., eye opening,
first estrus, vaginal opening) from Feng et al. (2017). The animal doses in the study, converted
to human equivalent doses (HEDs), were used in the BMD modeling; the data are available for
download in Health Assessment Workspace Collaborative (HAWC). BMD modeling was
conducted by experts in quantitative Benchmark Dose Software (BMDS) analysis and
interpretation. Links to the data and modeling output are included in Table F-l. The selected
point of departure (POD) (HED) listed in Table F-l represents the best-fitting model for each
endpoint; if the data were determined not to be amenable to BMD modeling, the
no-observed-adverse-effect level (NOAEL) or lowest-observed-adverse-effect level (LOAEL) is
listed. Figure F-l illustrates the doses examined and NOAEL, LOAEL, BMD, and benchmark
dose lower confidence limit (BMDL) values for the potential critical effects.
Table F-l. Candidate PODs for the Derivation of the Subchronic and Chronic RfDs for
PFBS (CASRN 375-73-5) and the Related Compound K+PFBS (CASRN 29420-49-3)
Endpoint/Reference
Species/Life Stage—Sex
Selected POD (HED)a
(mg/kg-d)
Kidney effects
Kidney histopathology—papillary epithelial tubular/ductal
hvDerolasia—Lieder et al. (2009a)
Rat—male
BMDLm = 0.489
Rat—female
BMDLio = 0.300
Kidney histopathology—papillary epithelial tubular/ductal
hvoerolasia—Lieder et al. (2009b)
Rat/Po—male
BMDL,o = 0.351
Rat/Po—female
BMDLm = 0.265
Kidney histopathology—papillary epithelial tubular/ductal
hvDerolasia—Lieder et al. (2009b)
Rat/Fi—male
BMDLm = 0.776
Rat/Fi—female
BMDLm = 0.478
Thyroid effects
Total T4—NTP (2019)
Rat—male
LOAEL = 0.34
Rat—female
BMDLisd = 0.037
Free T4—NTP (2019)
Rat—male
LOAEL = 0.34
Rat—female
BMDLisd = 0.027
Total Ti—Feng et al. (2017)
Mouse/Po—female
BMDLisd = 0.093
Free T4—Feng et al. (2017)
Mouse/Po—female
NOAEL = 0.21
TSH—Feng etal. (2017)
Mouse/Po—female
NOAEL = 0.21
Total T4 PND 1 (fetal n)h—Feng et al. (2017)
Mouse/Fi —female
NOAEL = 0.21
Total T» PND 1 (litter //)'—Feng et al. (2017)
Mouse/Fi—female
BMDLo.ssd = 0.095
(BMDLisd = 0.25)
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Table F-l. Candidate PODs for the Derivation of the Subchronic and Chronic RfDs for
PFBS (CASRN 375-73-5) and the Related Compound K+PFBS (CASRN 29420-49-3)
Endpoint/Reference
Species/Life Stage—Sex
Selected POD (HED)a
(mg/kg-d)
Total T4 PND 30—Feng et al. (2017)
Mouse/Fi —female
NOAEL = 0.21
Total T4 PND 60—Feng et al. (2017)
Mouse/Fi —female
NOAEL = 0.21
TSH PND 30—Feng et al. (2017)
Mouse/Fi —female
NOAEL = 0.21
Developmental effects
Eves opening (fetal n)h—Feng et al. (2017)
Mouse/Fi —female
NOAEL = 0.21
Eves ODcninu (litter n)h—Feng et al. (2017)
Mouse/Fi —female
BMDLo 5rd = 0.073
(BMDLisd = 0.16)
Vaginal opening (fetal n)h—Feng et al. (2017)
Mouse/Fi —female
BMDLo ssn — 0.15

(BMDLisd = 0.35)
Vaginal opening (litter n)h—Feng et al. (2017)
Mouse/Fi —female
BMDLo ssn — 0.094

(BMDLisd = 0.22)
First estrous (fetal n)h—Feng et al. (2017)
Mouse/Fi —female
NOAEL = 0.21
First estrous (litter n)h—Feng et al. (2017)
Mouse/Fi —female
NOAEL = 0.21
'Following U.S. EPA (2011b) guidance, animal doses from candidate principal studies were converted to HEDs
through the application of a DAF, where HED = dose x DAF. See Table 8 in the assessment for full details. Links
are to the HAWC BMDS session containing full modeling results for that endpoint.
' Fetal endpoints from Feng etal. (2017) were modeled alternatively using dose group sizes based either on total
number of fetuses or dams. Given that it appears that Feng et at (2017) did not use the litter as the statistical unit
of analysis, it is unclear if the study-reported standard errors pertain to litters or fetuses. Alternatively, modeling
fetal endpoints using litter n or fetal n provides two modeling results that bracket the "true" variance among all
fetuses in a dose group (i.e., using the fetal n will underestimate the true variance while using the litter n will
overestimate the true variance). Individual animal data were requested from study authors but were unable to be
obtained.
BMDL = benchmark dose lower confidence limit; BMDS = benchmark dose software; DAF = dosimetric
adjustment factor; HAWC = Health Assessment Workspace Collaborative; HED = human equivalent dose;
K+PFBS = potassium perfluorobutane sulfonate; LOAEL = lowest-observed-adverse-effect level;
NOAEL = no-observed-adverse-effect level; PFBS = perfluorobutane sulfonic acid; PND = postnatal day;
POD = point of departure; RfD = oral reference dose; SD = standard deviation; T4 = total thyroxine;
TSH = thyroid-stimulating hormone.
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EPA 690 R-2 J 00 IF
study name
experiment name
animal description
endpoint name
observation time
Lieder, 2009.1578546
90 Day Oral
Rat, Crl:Cd@(Sd)!gs Br Vaf/Plustm (cf)
Kidney, Hyperplasia, Papillary Tubular/Ductal Epithelium
90.0 days


Rat. Crl:Cd@(Sd)Igs Br Vaf/Plustm ( 9)
Kidney, Hyperplasia, Papillary Tubular/Ductal Epithelium
90.0 days
Lieder, 2009,1578545
2 Generation Oral
P0 Rat, Sprague-Dawley (cf)
Kidney, Hyperplasia, Papillary Tubular/Ductal Epithelium
None not reported


P0 Rat, Sprague-Dawley (9)
Kidney, Hyperplasia, Papillary Tubular/Ductal Epithelium
None not reported


PI Rat, Sprague-Dawley (cf)
Kidney, Hyperplasia, Papillary Tubular/Ductal Epithelium
120.0 days


F1 Rat, Sprague-Dawley (9)
Kidney, Hyperplasia, Papillary Tubular/Ductal Epithelium
120.0 days
NTP, 2018,4309741
28 Day Oral
Rat. Harlan Sprague-Dawley (cf)
Tetraiodothyronine (T4), Free
None not reported


Rat. Harlan Sprague-Dawley (9)
Tetraiodothyronine (T4), Free
None not reported


Rat, Harlan Sprague-Dawley (cf)
Tetraiodothyronine (T4). Total
None not reported


Rat, Harlan Sprague-Dawley (9)
Tetraiodothyronine (T4), Total
None not reported
Feng 2017,3856465
20 Day Oral Gestation
P0 Mouse. ICR (9)
Tetraiodothyronine (T4). Free
20.0 gestational day (GD)



Tetraiodothyronine (T4), Total
20.0 gestational day (GD)


Fl Mouse, ICR (9)
Tetraiodothyronine (T4), Total
Tetraiodothyronine (T4), Total - Litter N
1.0 post-natal day (PND)
30.0 post-natal day (PND)
60.0 post-natal day (PND)
1.0 post-natal day (PND)


P0 Mouse. ICR (9)
Thyroid Stimulating Hormone (TSH)
20.0 gestational day (GD)


Fl Mouse, ICR (9)
Thyroid Stimulating Hormone (TSH)
1.0 post-natal day (PND)
30.0 post-natal day (PND)
60.0 post-natal day (PND)



Eye Opening - Fetal N
12.0 post-natal day (PND)



Eye Opening - Litter N
12.0 post-natal day (PND)



First Estrous - Fetal N
24.0 post-natal day (PND)



First Estrous - Litter N
24.0 post-natal day (PND)



Vaginal Opening - Fetal N
24.0 post-natal day (PND)



Vaginal Patency - Litter N
24.0 post-natal day (PND)
I'KBS Candidate PODs for Kfl)s
o
o o
o o
o
-©—	®
—G	•	~
o BMD
O bmdl
G NOAEL
G LOAEL
• Doses
\—j Dose Range
o
-i—i 111111
c -
•©-
•e-
Dose (mg/kg/day)
Figure F-l. Candidate PODs for the Derivation of the Subchronic and Chronic RfDs for PFBS
(Click to see interactive data graphic)
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F.2. MODELING PROCEDURE FOR CONTINUOUS NONCANCER DATA
BMD modeling of continuous data was conducted on the HAWC website using the
U.S. Environmental Protection Agency's (U.S. EPA's) BMDS (Version 2.7). All continuous
models available within the software were fit using a benchmark response (BMR) of 1 standard
deviation (SD). For continuous data of effects in developing offspring, including thyroid
hormone changes, a BMR of 0.5 SD change from the control mean is used to account for effects
occurring in a sensitive life stage. A 1 SD BMR is also presented as the basis for model
comparison as directed in the U.S. EPA Benchmark Dose Technical Guidance (U.S. EPA. 2012).
An adequate fit is judged based on the x2 goodness-of-fit p-walue (p> 0.1), magnitude of the
scaled residuals in the vicinity of the BMR, and visual inspection of the model fit. In addition to
these three criteria forjudging adequacy of model fit, a determination is made as to whether the
variance across dose groups is homogeneous. If a homogeneous variance model is deemed
appropriate based on the statistical test provided by BMDS (i.e., Test 2), the final BMD results
are estimated from a homogeneous variance model. If the test for homogeneity of variance is
rejected (p < 0.1), the model is run again while modeling the variance as a power function of the
mean to account for this nonhomogeneous variance. If this nonhomogeneous variance model
does not adequately fit the data (i.e., Test 3;p<0. 1), the data set is considered unsuitable for
BMD modeling. In cases in which a model with # parameters = # dose-groups was fit to the data
set, all parameters were estimated, and no p-w alue was calculated, that model was not considered
for estimating a POD unless no other model provided adequate fit. Among all models providing
adequate fit, the BMDL from the model with the lowest Akaike's information criterion (AIC)
was selected as a potential POD when BMDL values were sufficiently close (within threefold).
Otherwise, the lowest BMDL was selected as a potential POD from which to derive the oral
reference dose/inhalation reference concentration (RfD/RfC).
F.2.1 Modeling Predictions for Serum Total T4 in PND 1 Female Offspring (litter n)
The modeling results for total T4 in PND 1 female offspring (litter n) exposed Gestation
Days (GDs) 1-20 are shown in Table F-2. The Exponential 4 model (see Figure F-2) was
selected given appropriate fit to the data and that the BMDL values differed by greater than
threefold. The output for the U.S. EPA's BMDS model run is also provided below.
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EPA 690 R-21 00IF
Table F-2. Modeling Results for Total T4 in PND 1 Female Offspring (Litter n)
Exposed GDs l-20a
Model
Global
/7-Value
AIC
BMDo.ssd
(HED)
(mg/kg-d)
BMDLo.ssd
(HED)
(mg/kg-d)
BMDisd
(HED)
(mg/kg-d)
BMDLisd
(HED)
(mg/kg-d)
Residual of
Interest
Linear
0.5652
-4.74898
0.7778
0.5120
1.5557
1.0241
0.348
Polynomial
0.5652
-4.74898
0.7778
0.5120
1.5557
1.0241
0.348
Power
0.5652
-4.74898
0.7778
0.5120
1.5557
1.0241
0.348
Hill
-999
-1.89
0.368
0.0704
0.8677
0.2294
-6.01 x 10-7
Exponential-M2
0.77
-5.3672
0.5546
0.3017
1.2555
0.6694
-0.5752
Exponential-M3
0.77
-5.3672
0.5546
0.3017
1.2555
0.6694
-0.5752
Exponential-M4b
0.8583
-3.8581
0.3346
0.0951
0.8708
0.2498
-0.08305
Exponential-M5
-999
-1.89
0.3807
0.0958
0.8669
0.2517
-4.356 x 10~7
aFeng et al. (2017).
bSelected model. Exponential 4 model was selected given appropriate fit to the data and that the BMDL values
differed by greater than threefold. The Hill and Exponential 5 models were not selected because they did not
return a /?-valuc.
AIC = Akaike's information criterion; BMD = maximum likelihood estimate of the exposure concentration
associated with the selected BMR; BMDL = 95% lower confidence limit on the BMD (subscripts denote BMR:
i.e., 0.5 SD = exposure concentration associated with 0.5 SD change from the control mean); BMR = benchmark
response; GD = gestation day; HED = human equivalent dose; PND = postnatal day; SD = standard deviation;
T4 = thyroxine.
Tetraiodothyronine (T4), Total - Litter N
O)
CD
(/)
C
o
Q_
C/)
CD
cr
O Doses in Study
O LOAEL
O NOAEL
• Exponential-M4
0.5 1.0 1.5 2.0
Dose (mg/kg-day HED)
Figure F-2. Exponential (Model 4) for Total T4 in PND 1 Female Offspring (Litter n)
Exposed GDs 1-20 (Feng et al., 2017)
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Exponential Model. (Version: 1.11; Date: 03/14/2017)
Input Data File: C:\Windows\TEMP\bmds-dfile-k4vsthrz.(d)
Gnuplot Plotting File:
Mon Aug 17 15:16:06 2020

BMDS_Model_Run

The form of the response function by Model:
Model 2:
Model 3:
Model 4:
Model 5:
Y[dose]
Y[dose]
Y[dose]
Y[dose]
a * exp{sign *	b * dose}
a * expjsign *	(b * dose)Ad}
a * [c-(c-l)	*	exp{-b * dose}]
a * [c-(c-l)	*	exp{-(b * dose)
d}]
Note: Y[dose] is the median response for exposure = dose;
sign = +1 for increasing trend in data;
sign = -1 for decreasing trend.
Model 2 is nested within Models 3 and 4.
Model 3 is nested within Model 5.
Model 4 is nested within Model 5.
Dependent variable = Response
Independent variable = Dose
Data are assumed to be distributed: normally
Variance Model: exp(lnalpha +rho *ln(Y[dose]))
rho is set to 0.
A constant variance model is fit.
Total number of dose groups = 4
Total number of records with missing values = 0
Maximum number of iterations = 500
Relative Function Convergence has been set to: le-008
Parameter Convergence has been set to: le-008
MLE solution provided: Exact
Initial Parameter Values
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EPA/690/R-21/001F
Variable	Model 4
lnalpha	-1.29725
c	0.434618
Parameter Estimates
Variable	Model 4	Std. Err.
C	0.417162	0.225239
NC = No Convergence
Table of Stats From Input Data
Dose	N	Obs Mean Obs Std Dev
0	10	1.44	0.329
0.21	10	1.3	0.657
0.86	10	0.92	0.493
2.14	10	0.69	0.657
Estimated Values of Interest
Dose	Est Mean	Est Std Scaled Residual
0	1.453	0.523	-0.07759
0.21	1.278	0.523	0.1354
0.86	0.9337	0.523	-0.08305
2.14	0.6858	0.523	0.02529
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Other models for which likelihoods are calculated:
Model Al:	Yij = Mu(i) + e(ij)
Var{e(ij)} = SigmaA2
Model A2:	Yij = Mu(i) + e(ij)
Var{e(ij)} = Sigma(i)A2
Model A3:	Yij = Mu(i) + e(ij)
Var{e(ij)} = exp(lalpha + log(mean(i)) * rho)
Model R:	Yij = Mu + e(i)
Var{e(ij)} = SigmaA2
Likelihoods of Interest
Model	Log(likelihood)	DF	AIC
Al	5.944999	5	-1.889998
A2	8.698072	8	-1.396144
A3	5.944999	5	-1.889998
R	0.3138778	2	3.372244
4	5.929054	4	-3.858109
Additive constant for all log-likelihoods = -36.76. This constant added to
the
above values gives the log-likelihood including the term that does not
depend on the model parameters.
Explanation of Tests
Test 1: Does response and/or variances differ among Dose levels? (A2 vs. R)
Test 2: Are Variances Homogeneous? (A2 vs. Al)
Test 3: Are variances adequately modeled? (A2 vs. A3)
Test 6a: Does Model 4 fit the data? (A3 vs 4)
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Tests of Interest
Test
Test 1
Test 2
Test 3
Test 6a
-2*log(Likelihood Ratio)
16.77
5.506
5.506
0.03189
6
3
3
1
p-value
0.01017
0.1383
0.1383
0.8583
The p-value for Test 1 is less than .05. There appears to be a
difference between response and/or variances among the dose
levels, it seems appropriate to model the data.
The p-value for Test 2 is greater than .1. A homogeneous
variance model appears to be appropriate here.
The p-value for Test 3 is greater than .1. The modeled
variance
appears to be appropriate here.
The p-value for Test 6a is greater than .1. Model 4 seems
to adequately describe the data.
Benchmark Dose Computations:
Specified Effect
Risk Type
Confidence Level = 0.950000
1.000000
= Estimated standard deviations from control
BMD =
0.87078
BMDL = 0.249811
BMDU =
21400
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Exponential Model. (Version: 1.11; Date: 03/14/2017)
Input Data File: C:\Uindows\TEMP\bmds-dfile-171ffb4f.(d)
Gnuplot Plotting File:
Mon Aug 17 15:16:07 2020
BMDS_Model_Run
The form of the response function by Model:
Model 2: Y[dose] = a * exp{sign * b * dose}
Model 3: Y[dose] = a * exp{sign * (b * dose)Ad}
Model 4: Y[dose] = a * [c-(c-1) * exp{-b * dose}]
Model 5: Y[dose] = a * [c-(c-1) * exp{-(b * dose)Ad}]
Note: Y[dose] is the median response for exposure = dose;
sign = +1 for increasing trend in data;
sign = -1 for decreasing trend.
Model 2 is nested within Models 3 and 4.
Model 3 is nested within Model 5.
Model 4 is nested within Model 5.
Dependent variable = Response
Independent variable = Dose
Data are assumed to be distributed: normally
Variance Model: exp(lnalpha +rho *ln(Y[dose]))
rho is set to 0.
A constant variance model is fit.
Total number of dose groups = 4
Total number of records with missing values = 0
Maximum number of iterations = 500
Relative Function Convergence has been set to: le-008
Parameter Convergence has been set to: le-008
MLE solution provided: Exact
Initial Parameter Values
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Variable	Model 4
lnalpha	-1.29725
c	0.434618
Parameter Estimates
Variable	Model 4	Std. Err.
C	0.417162	0.225239
NC = No Convergence
Table of Stats From Input Data
Dose	N	Obs Mean Obs Std Dev
0	10	1.44	0.329
0.21	10	1.3	0.657
0.86	10	0.92	0.493
2.14	10	0.69	0.657
Estimated Values of Interest
Dose	Est Mean	Est Std Scaled Residual
0	1.453	0.523	-0.07759
0.21	1.278	0.523	0.1354
0.86	0.9337	0.523	-0.08305
2.14	0.6858	0.523	0.02529
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Other models for which likelihoods are calculated:
Model Al:	Yij = Mu(i) + e(ij)
Var{e(ij)} = SigmaA2
Model A2:	Yij = Mu(i) + e(ij)
Var{e(ij)} = Sigma(i)A2
Model A3:	Yij = Mu(i) + e(ij)
Var{e(ij)} = exp(lalpha + log(mean(i)) * rho)
Model R:	Yij = Mu + e(i)
Var{e(ij)} = SigmaA2
Likelihoods of Interest
Model	Log(likelihood)	DF	AIC
Al	5.944999	5	-1.889998
A2	8.698072	8	-1.396144
A3	5.944999	5	-1.889998
R	0.3138778	2	3.372244
4	5.929054	4	-3.858109
Additive constant for all log-likelihoods = -36.76. This constant added to
the
above values gives the log-likelihood including the term that does not
depend on the model parameters.
Explanation of Tests
Test 1: Does response and/or variances differ among Dose levels? (A2 vs. R)
Test 2: Are Variances Homogeneous? (A2 vs. Al)
Test 3: Are variances adequately modeled? (A2 vs. A3)
Test 6a: Does Model 4 fit the data? (A3 vs 4)
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EPA/690/R-21/001F
Tests of Interest
Test
Test 1
Test 2
Test 3
Test 6a
-2*log(Likelihood Ratio)
16.77
5.506
5.506
0.03189
p-value
6
3
3
1
0.01017
0.1383
0.1383
0.8583
The p-value for Test 1 is less than .05. There appears to be a
difference between response and/or variances among the dose
levels, it seems appropriate to model the data.
The p-value for Test 2 is greater than .1. A homogeneous
variance model appears to be appropriate here.
The p-value for Test 3 is greater than .1. The modeled
variance appears to be appropriate here.
The p-value for Test 6a is greater than .1. Model 4 seems
to adequately describe the data.
Benchmark Dose Computations:
Specified Effect = 0.500000
Risk Type = Estimated standard deviations from control
Confidence Level = 0.950000
BMD
0.33455
BMDL = 0.0950923
BMDU =
1.22544
F.3 MODELING PROCEDURE FOR DICHOTOMOUS NONCANCER DATA
BMD modeling of dichotomous noncancer data (see Figure F-l) was conducted on the
HAWC website using the U.S. EPA's BMDS Version 2.7. For these data, the Gamma, Logistic,
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EPA/690/R-21/001F
Log-Logistic, Log-Probit, Multistage, Probit, and Weibull dichotomous models available within
the software were fit using a BMR of 10% extra risk. The Multistage model is run for all
polynomial degrees up to n - 2, where n is the number of dose groups including control.
Adequacy of model fit was judged based on the x2 goodness-of-fit p-value (p > 0.1), scaled
residuals at the data point (except the control) closest to the predefined BMR (absolute
value < 2.0), and visual inspection of the model fit. In the cases where no best model was found
to fit to the data, use of a reduced data set without the high-dose group was further attempted for
modeling and the result was presented along with that of the full data set. In cases in which a
model with # parameters = # dose-groups was fit to the data set, all parameters were estimated,
and no p-v alue was calculated, that model was not considered for estimating a POD unless no
other model provided adequate fit. Among all models providing adequate fit, the BMDL from
the model with the lowest AIC was selected as a potential POD when BMDL values were
sufficiently close (within threefold) (see Table F-l). Otherwise, the lowest BMDL was selected
as a potential POD.
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APPENDIX G. QUALITY ASSURANCE
U.S. EPA has an agency-wide quality assurance (QA) policy, and that policy is outlined
in the EPA Quality Manual for Environmental Programs (see CIO 2105-P-01-0) and follows the
specifications outlined in U.S. EPA Order CIO 2105.0. The goal of the QA policy is to assure
that environmental data used to support Agency decisions are of adequate quality and usability
for their intended purpose.
As required by CIO 2105.0. ORD maintains a Quality Management Program, which is
documented in an internal Quality Management Plan (QMP). The latest version was developed
in 2013 using the Guidance for Developing Quality Systems for Environmental Programs
(QA/G-1). An NCEA-specific QMP was also developed in 2013 as an appendix to the ORD
QMP. Quality assurance for products developed within CPHEA is managed under the ORD
QMP and applicable appendices.
This assessment has been designated as High Profile and is classified as QA Category A.
Category A designations require reporting of all critical QA activities, including audits.
Another requirement of the Agency quality system includes the use of project-specific
planning documents referred to as Quality Assurance Project Plans (QAPPs) that describe how
specific data collection efforts will be planned, implemented, and assessed. Specific
management of quality assurance in this assessment is documented in an Umbrella Quality
Assurance Project Plan, which was developed using the U.S. EPA Guidance for Quality
Assurance Project Plans (OA/G-5). The latest approved version of the QAPP is dated September
2019. During assessment development, additional QAPPs may be applied for quality assurance
management. They include:
Title
Document Number
Date
Program Quality Assurance Project Plan (PQAPP) for the
Provisional Peer-Reviewed Toxicity Values (PPRTVs) and Related
Assessments/Documents
L-CP AD-0032718-QP
October 2015 (last
updated 2020)
Umbrella Quality Assurance Project Plan for NCEA PFAS Toxicity
Assessments
B-IO-0031652-QP-1-2
July 2018 (last
updated September
2019)
Quality Assurance Project Plan (QAPP) for Enhancements to
Benchmark Dose Software (BMDS)
B-003742-QP-1-0
July 2019
During assessment development, this project underwent quality audit:
Date
Type of Audit
Major Findings
Actions Taken
September 18, 2020
Technical System Audit
None
None
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During assessment development, the assessment was subjected to external reviews by
individual letters from expert peer reviewers and by other federal agency partners including the
Executive Offices of the President. Peer-review reports during these review steps are available
at https://www.epa.gov/pfas/leam-about-human-health-toxicitv-assessment-pfbs. In addition, the
assessment underwent public comment from November 21, 2018 to January 22, 2019. The
public comments are available in the Docket ID No. EPA-HQ-OW-2018-0614. Prior to release,
the final draft assessment was submitted to management and QA clearance. During this step the
CPHEA QA director and QA managers review the project QA documentation and ensure
U.S. EPA QA requirements have been met.
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