JHH f United States
Environmental Protection
R mAgency
EP A-822-P-23 -004
PUBLIC REVIEW DRAFT
Maximum Contaminant Level Goal (MCLG) Summary
Document for a Mixture of Four Per- and Polyfluoroalkyl
Substances (PFAS):
HFPO-DA and its Ammonium Salt (also known as GenX
Chemicals), PFBS, PFNA, and PFHxS
CASRN 13252-13-6 and 62037-80-3 (HFPO-DA)
CASRN 375-73-5 and 29420-49-3 (PFBS)
CASRN 375-95-1 (PFNA)
CASRN 355-46-4 (PFHxS)
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Maximum Contaminant Level Goal (MCLG) Summary Document for a
Mixture of Four Per- and Polyfluoroalkyl Substances (PFAS):
HFPO-DA and its Ammonium Salt (also known as GenX Chemicals), PFBS,
PFNA, and PFHxS
CASRNs 13252-13-6 and 62037-80-3 (HFPO-DA)
CASRNs 375-73-5 and 29420-49-3 (PFBS)
CASRN 375-95-1 (PFNA)
CASRN 355-46-4 (PFHxS)
Prepared by:
U.S. Environmental Protection Agency
Office of Water (4304T)
Office of Science and Technology
Health and Ecological Criteria Division
Washington, DC 20460
EPA Document Number: EPA-822-P-23-004
March 2023
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Notices
This document has been reviewed in accordance with EPA policy and approved for publication.
This document provides a summary of information used to develop the proposed MCLG for the
mixture of HFPO-DA and its ammonium salt (also known as GenX chemicals)1, PFBS, PFNA,
and PFHxS.
Mention of trade names or commercial products does not constitute endorsement or
recommendation for use.
:EPA notes that the chemical HFPO-DA is used in a processing aid technology developed by DuPont to make fluoropolymers
without using PFOA. The chemicals associated with this process are commonly known as GenX Chemicals and the term is often
used interchangeably for HFPO-DA along with its ammonium salt.
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Authors, Contributors, and Reviewers
Prepared by
Carlye Austin, PhD
Czarina Cooper, MPH
Colleen Flaherty, MS
Contributors
Brandi Echols, PhD
Alexis Lan, MPH
EPA Technical Review
Office of Chemical Safety and Pollution Prevention
Office of Children's Health Protection
Office of Land and Emergency Management
Office of Research and Development
Executive Direction
Elizabeth Behl
Eric Burneson, P.E.
Formatting By
Tetra Tech, Inc
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Contents
Abbreviations and Acronyms iv
1.0 Introduction and Background 1
1.1 Purpose 1
1.2 Co-Occurrence of PFAS in Drinking Water 1
1.3 Dose Additivity for PFAS 2
1.4 Mixture Approaches Considered 3
1.5 Overview of Mixture Hazard Index (HI) MCLG Approach 4
2.0 Health-Based Water Concentrations 8
2.1 HFPO-DA 8
2.1.1 Toxicity 8
2.1.2 Exposure Factor 9
2.1.3 Relative Source Contribution 9
2.1.4 Derivation of HFPO-DA HBWC 9
2.2 PFBS 10
2.2.1 Toxicity 10
2.2.2 Exposure Factor 11
2.2.3 Relative Source Contribution 11
2.2.4 Derivation of PFBS HBWC 11
2.3 PFNA 12
2.3.1 Toxicity 13
2.3.2 Exposure Factor 13
2.3.3 Relative Source Contribution 14
2.3.4 Derivation of PFNA HBWC 14
2.4 PFHxS 15
2.4.1 Toxicity 15
2.4.2 Exposure Factor 16
2.4.3 Relative Source Contribution 16
2.4.4 Derivation of PFHxS HBWC 16
3.0 Derivation of PFAS Mixture Hazard Index MCLG 18
References 19
APPENDIX A. PFNA: Summary of Occurrence in Water and Detailed Relative Source
Contribution 25
Occurrence in Water 25
Drinking Water 25
Groundwater 25
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Surface Water 26
RSC for PFNA 26
Literature Search and Screening 26
Additional Screening 28
Summary of Potential Sources of PFNA Exposure 29
Dietary Sources 29
Food Contact Materials 43
Consumer Products 43
Indoor Dust 46
Air 53
Soil 53
Sediment 62
Biomonitoring in the U.S. Population 62
Recommended RSC 62
References 62
APPENDIX B. PFHxS: Summary of Occurrence in Water and Detailed Relative Source
Contribution 71
Occurrence in Water 71
Drinking Water 71
Groundwater 71
Surface Water 72
RSC for PFHxS 72
Literature Search and Screening 72
Additional Screening 74
Summary of Potential Sources of PFHxS Exposure 75
Dietary Sources 75
Consumer Products 89
Indoor Dust 91
Air 97
Soil 97
Sediment 106
Biomonitoring in the U.S. Population 106
Recommended RSC 106
References 106
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Tables
Table 1. HFPO-DA HBWC - Input Parameters and Value 10
Table 2. PFBS HBWC - Input Parameters and Value 12
Table 3. PFNA HBWC - Input Parameters and Value 14
Table 4. PFHxS HBWC - Input Parameters and Value 17
Table A-l. Populations, Exposures, Comparators, and Outcomes (PECO) Criteria 28
Table A-2. Summary of PFNA Data in Seafood 32
Table A-3. Summary of PFNA Data in Other Food 38
Table A-4. Summary of PFNA Consumer Product Data 45
Table A-5. Summary of PFNA Indoor Dust Data 49
Table A-6. Summary of PFNA Data in Soil 56
Table B-l. Populations, Exposures, Comparators, and Outcomes (PECO) Criteria 74
Table B-2. Summary of PFHxS Data in Seafood 77
Table B-3. Summary of PFHxS Data in Other Food 83
Table B-4. Summary of PFHxS Consumer Product Data 90
Table B-5. Summary of PFHxS Indoor Dust Data 94
Table B-6. Summary of PFHxS Data in Soil 100
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Abbreviations and Acronyms
AFFF
aqueous film-forming
foam
AT SDR
Agency for Toxic
Substances and Disease
Registry
BDL
below the detection limit
BMD
benchmark dose
CDC
Centers for Disease
Control and Prevention
CTEPP
Children's Total
Exposure to Persistent
Pesticides and Other
Persistent Organic
Pollutants
DA
dose addition
DWI-BW
body weight-adjusted
drinking water intake
E
duration-relevant
exposure
EPA
U.S. Environmental
Protection Agency
FAO
Food and Agriculture
Organization Area
FDA
U.S. Food and Drug
Administration
FIFRA
Federal Insecticide,
Fungicide, and
Rodenticide Act
GCA
groundwater
contamination area
GenX chemicals
hexafluoropropy 1 ene
oxide (HFPO) dimer acid
and HFPO dimer acid
ammonium salt
GD
gestational day
HA
health advisory
HBWC
health-based water
concentration
HED
human equivalent dose
HDPE
high density polyethylene
HFPO
hexafluoropropy 1 ene
oxide
HI
hazard index
HQ
hazard quotient
IRIS
Integrated Risk
Information System
L/kg/day
liters per kilogram body
weight per day
LOAEL
lowest-observed-adverse-
effect level
LOD
limit of detection
LOQ
limit of quantitation
MAMA
Methods Advancement
for Milk Analysis
MCLG
Maximum Contaminant
Level Goal
MDL
method detection limit
MF
modifying factor
mg/kg/day
milligrams per kilogram
body weight per day
mg/L
milligrams per liter
MOA
mode of action
MRL
minimal risk level
MW
molecular weight
ng/L
nanograms per liter
NHANES
National Health and
Nutrition Examination
Survey
NO A A
National Oceanic and
Atmospheric
Administration
NOAEL
no-ob served-adverse-
effect level
NPDWR
National Primary
Drinking Water
Regulation
NRSA
National Rivers and
Streams Assessment
NTP
National Toxicology
Program
PECO
Population, Exposure,
Comparator, and
Outcome
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PFAS
per- and polyfluoroalkyl
substances
PFBS
perfluorobutanesulfonic
acid
PFHpA
perfluoroheptanoic acid
PFHxA
perfluorohexanoic acid
PFHxS
perfluorohexanesulfonic
acid
PFNA
perfluorononanoic acid
PFOA
perfluorooctanoic acid
PFOS
perfluorooctanesulfonic
acid
PND
postnatal day
POD
point of departure
ppt
parts per trillion
PSA
prostate-specific antigen
PVDF
polyvinylidene fluoride
PWS
public water system
RfD
reference dose
RfV
reference value
RSC
relative source
contribution
MARCH 2023
SAB
Science Advisory Board
SDWA
Safe Drinking Water Act
UC MR
Unregulated Contaminant
Monitoring Rule
UCMR3
third Unregulated
Contaminant Monitoring
Rule
UF
uncertainty factor
UFa
interspecies uncertainty
factor
UFd
database uncertainty
factor
UFh
human interindividual
variability uncertainty
factor
UFs
extrapolation from
sub chronic to chronic
exposure duration
uncertainty factor
^g/L
micrograms per liter
wos
Web of Science
WWTP
wastewater treatment
plant
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1.0 Introduction and Background
1.1 Purpose
Section 1412(a)(3) of the Safe Drinking Water Act (SDWA) requires the Administrator of the
U.S. Environmental Protection Agency (EPA) to propose a Maximum Contaminant Level Goal
(MCLG) simultaneously with the National Primary Drinking Water Regulation (NPDWR). The
MCLG is set, as defined in Section 1412(b)(4)(A), at "the level at which no known or anticipated
adverse effects on the health of persons occur and which allows an adequate margin of safety."
Consistent with SDWA 1412(b)(3)(C)(i)(V), in developing the MCLG, EPA considers "the
effects of the contaminant on the general population and on groups within the general population
such as infants, children, pregnant women, the elderly, individuals with a history of serious
illness, or other subpopulations that are identified as likely to be at greater risk of adverse health
effects due to exposure to contaminants in drinking water than the general population." Other
factors considered in determining MCLGs include health effects data on drinking water
contaminants and potential sources of exposure other than drinking water. MCLGs are not
regulatory levels and are not enforceable.
The purpose of this document is to provide a summary of the health effects and exposure
information and analyses, and to describe the derivation of EPA's proposed MCLG for a mixture
of the following per- and polyfluoroalkyl substances (PFAS): hexafluoropropylene oxide
(HFPO) dimer acid and its ammonium salt (also known as GenX chemicals)2, perfluorobutane
sulfonic acid and its related compound potassium perfluorobutane sulfonate (PFBS),
perfluorononanoic acid (PFNA), and perfluorohexanesulfonic acid (PFHxS).3 The PFAS mixture
MCLG is based on a hazard index (HI) approach, a commonly used component-based mixture
risk assessment method (see Section I.D and EPA, 2022c). This document is not intended to be
an exhaustive description of all health effects or modeled endpoints (i.e., human health toxicity
assessment) nor a drinking water health advisory (HA); rather, this document summarizes key
elements (e.g., reference doses (RfDs) from recently published, peer-reviewed, publicly available
assessments for HFPO-DA (EPA, 2021a; 2022a), PFBS (EPA, 2021b; 2022b), PFNA (ATSDR,
2021), and PFHxS (ATSDR, 2021) that EPA used to develop health-based water concentrations
(HBWCs) that inform the proposed MCLG for HFPO-DA, PFBS, PFNA, and PFHxS.
1.2 Co-Occurrence of PFAS in Drinking Water
Improved analytical monitoring and detection methods have enabled detection of the co-
occurrence of multiple PFAS in drinking water, ambient surface waters, aquatic organisms,
biosolids (sewage sludge), and other environmental media (EPA, 2022a,b; 2023a,b,c,d,e,f).
PFOA and PFOS have historically been target analytes, and the focus of many environmental
monitoring studies. More recent monitoring studies, however, have begun to focus on additional
PFAS via advanced analytical instruments/methods and non-targeted analysis (De Silva et al.,
2020; McCord and Strynar, 2019; McCord et al., 2020).
2EPA notes that the chemical HFPO-DA is used in a processing aid technology developed by DuPont to make fluoropolymers
without using PFOA. The chemicals associated with this process are commonly known as GenX Chemicals and the term is often
used interchangeably for HFPO-DA along with its ammonium salt.
3 Note: EPA is also proposing individual MCLGs for two PFAS (perfluorooctanoic acid (PFOA) and
perfluorooctanesulfonic acid (PFOS); see EPA 2023a,b,c,d-
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EPA uses the Unregulated Contaminant Monitoring Rule (UCMR) to collect occurrence data for
contaminants that are suspected to be present in drinking water and do not have health-based
standards under SDWA. Between 2013 and 2015, EPA's third UCMR (UCMR 3) required all
large public water systems (PWSs) (each serving more than 10,000 people) and a statistically
representative national sample of 800 small PWSs (each serving 10,000 people or fewer) to
monitor for 30 unregulated contaminants in drinking water, including six PFAS: PFOS, PFOA,
PFNA, PFHxS, perfluoroheptanoic acid (PFHpA), and PFBS. UCMR 3 data demonstrated that
two or more of those six PFAS co-occurred in 48% (285/599) of sampling events with PFAS
detected, and PFOA and PFOS co-occurred in 27% (163/599) of sampling events with two or
more PFAS detected (EPA, 2019b). EPA found that 4% of PWSs reported results for which one
or more of the six UCMR 3 PFAS were measured at or above their respective minimum
reporting levels.4 In addition to the UCMR 3 data collection, many states have undertaken more
recent efforts to monitor for PFAS in both source and finished drinking water using newer
analytical methods and reflecting lower reporting limits than those in UCMR 3. These results
show continued PFAS occurrence and co-occurrence in multiple geographic locations. These
data also show certain PFAS, including PFOS, PFOA, PFNA, PFHxS, and PFBS, at lower
concentrations and significantly greater frequencies than were measured under UCMR 3.
Additionally, these state monitoring data include results for HFPO-DA (which were not included
in the suite of PFAS analyzed in UCMR 3) and demonstrate HFPO-DA (and co-occurrence with
other PFAS) in drinking water in at least nine states (EPA, 2023f). In 2023-2025, UCMR 5 will
collect new monitoring data on 29 PFAS including HFPO dimer acid, PFBS, PFNA, and PFHxS,
which will help increase EPA's understanding of PFAS occurrence and co-occurrence in
drinking water.
Further discussion of the occurrence of HFPO-DA, PFBS, PFNA, and PFHxS in drinking water
(and other environmental media) can be found in EPA's occurrence Technical Support
Document (EPA 2023f).
1.3 Dose Additivity for PFAS
PFAS, including HFPO-DA, PFBS, PFNA, and PFHxS, disrupt signaling of multiple biological
pathways resulting in common adverse effects on several biological systems and functions,
including thyroid hormone levels, lipid synthesis and metabolism, development, and immune and
liver function (ATSDR, 2021; EFSA, 2018, 2020; EPA, 2022c). EPA has developed a
Framework for Estimating Noncancer Health Risks Associated with Mixtures of Per- and
Polyfluoroalkyl Substances (PFAS) (hereafter called "PFAS Mixtures Framework") (EPA,
2022c), based on existing EPA mixtures guidelines and guidance (EPA, 1986, 2000a). The PFAS
Mixtures Framework describes a flexible, data-driven approach that facilitates practical
component-based mixtures evaluation of two or more PFAS based on dose additivity. All of the
approaches described in the PFAS Mixtures Framework, including the HI approach (Section III),
involve integrating dose-response metrics that have been scaled based on the potency of each
PFAS in the mixture.
4 The 4% figure is based on 198 PWSs reporting measurable PFAS results for one or more sampling events from
one or more of their sampling locations. Those 198 PWSs serve an estimated total population of approximately 16
million (EPA, 2019b,c).
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Because PFAS are an emerging chemical class of note for toxicological evaluations and human
health risk assessment, mode of action (MOA) data may be limited or not available for many
PFAS. As such, the component-based approaches for assessing risks of PFAS mixtures are
focused on evaluation of similarity of toxicological endpoint/effect rather than similarity in
MO A, consistent with EPA mixtures guidance (EPA, 2000a). Precedents of prior research
conducted on mixtures of various chemical classes with disparate molecular initiating events, but
common key events5 or adverse outcomes, support the use of dose additive models for
estimating mixture-based risks (EPA, 2022c). In particular, Conley et al., 2022 recently
investigated in vivo effects in maternal rats and offspring from combined exposure to PFOA and
PFOS during gestation and early lactation. The study included a series of experiments designed
to characterize dose response curves across multiple endpoints for PFOA and PFOS individually,
followed by a mixture study of the two chemicals combined. The mixture experiment was
designed to test for shifts in the PFOA dose response curves from combined exposure to a fixed
dose of PFOS and to compare dose additivity and response additivity model predictions. For
nearly all endpoints amenable to mixture model analyses, the dose addition equation produced
equivalent or better estimates of observed data than response addition (a detailed discussion of
PFAS mixtures research, including dose additive and response additive models can be found in
EPA, 2022c). Thus, in the absence of detailed characterization of molecular mechanisms for
most PFAS, it is considered a reasonable health-protective assumption that PFAS which can be
demonstrated to share one or more key events or adverse outcomes will produce dose-additive
effects from co-exposure (EPA, 2022c). EPA received a generally favorable review from EPA's
Science Advisory Board (SAB) (EPA SAB, 2022) for its development of approaches based on
dose additivity, including the HI approach, to evaluate and manage risk from PFAS mixtures in
drinking water and other environmental media.6 For a detailed description of the evidence
supporting dose additivity for PFAS, see EPA (2022c).
1.4 Mixture Approaches Considered
There has been a lot of work evaluating parameters that best inform the combining of PFAS
components identified in environmental matrices into mixtures analyses. Indeed, there is
currently no consensus on whether or how PFAS should be combined for risk assessment
purposes. EPA considered several approaches to account for dose additive noncancer effects
associated with HPFO-DA, PFBS, PFNA, and PFHxS in mixtures. PFAS can affect multiple
human health endpoints and differ in their impact (i.e., potency of effect) on target
organs/systems. For example, one PFAS may be most toxic to the liver, and another may be most
toxic to the thyroid but both chemicals affect the liver and the thyroid. Other chemicals regulated
as groups operate through a common mode of action and predominately affect one human health
endpoint. This supports a flexible data-driven approach that facilitates the evaluation of multiple
health endpoints, such as the hazard index.
EPA considered the two main types of HI approaches: 1) the general HI which allows for
component chemicals in the mixture to have different health effects or endpoints as the basis for
the component chemical reference values (e.g., RfDs), and 2) the target-organ specific HI which
5 "Key event" is an empirically observable precursor step that is itself a necessary element of the mode of action or
is a biologically based marker for such an element (EPA, 2005).
6 "The SAB supports dose additivity based on a common outcome, instead of a common mode of action as a health
protective default assumption and does not propose another default approach." (EPA SAB, 2022)
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relies on reference values based on the same organ or organ system (e.g., liver-, thyroid-, or
developmental-specific). The general HI approach is based on the overall RfD which is
protective of all effects, and thus is a more health protective indicator of risk. The target-organ
specific HI approach produces a less health protective estimate of risk than the general HI when
a contaminant impacts multiple organs because the range of potential effects has been scoped to
a specific target organ, which may be one of the less potent effects or for which there may be
significant currently unquantified effects. Additionally, a target-organ specific HI approach relies
on toxicity values aggregated by the "same" target organ endpoint/effect, and the absence of
information about a specific endpoint may result the contaminant not being adequately
considered in a target-organ specific approach, and thus, underestimating potential health risk. A
target-organ specific HI can only be performed for those PFAS for which a health effect specific
RfD is calculated. For example, for some PFAS a given health effect might be poorly
characterized or not studied at all, or, as a function of dose may be one of the less(er) potent
effects in the profile of toxicity for that particular PFAS. Another limitation is that so many
PFAS lack human epidemiological or experimental animal hazard and dose-response information
across a broad(er) effect range thus limiting derivation of target-organ specific values. A similar,
effect/endpoint-specific method called the relative potency factor (RPF) approach, which
represents the relative difference in potency of an effect/endpoint between an index chemical and
other members of the mixture, was also considered. (Further background on all of these
approaches, plus illustrative examples, and a discussion of the advantages and challenges
associated with each approach can be found in Section 5 and 6 in EPA 2022c).
EPA also considered setting individual MCLGs instead of and in addition to using a mixtures-
based approach for HFPO-DA, PFBS, PFNA, and/or PFHxS in mixtures. EPA ultimately
selected the general HI approach for establishing an MCLG for these four PFAS, as described in
greater detail below, because it provides the most health protective endpoint for multiple PFAS
in a mixture to ensure there would be no known or anticipated adverse effects on the health of
persons. EPA also considered a target- specific HI or relative potency factor approach but,
because of information gaps, EPA may not be able to ensure that the MCLG is sufficiently health
protective. If the Agency only established an individual MCLG, the Agency would not provide
any protection against dose-additivity from regulated co-occuring PFAS (see Rule Preamble for
additional discussion).
1.5 Overview of Mixture Hazard Index (HI) MCLG Approach
For chemicals exhibiting a noncancer threshold for toxic effects (Category II or III; e.g., see
EPA, 1985; 1991) and nonlinear carcinogens (e.g., see EPA, 2006), EPA establishes the MCLG
based on a toxicity value, typically an RfD, but similar toxicity values may also be used when
they represent the best available science (e.g., Agency for Toxic Substances and Disease
Registry (ATSDR) Minimal Risk Level7 (MRL)). The MCLG is designed to be protective of
effects over a lifetime of exposure with an adequate margin of safety, including for sensitive
populations and life stages consistent with SDWA 1412(b)(3)(C)(i)(V) and 1412(b)(4)(A). The
calculation of a MCLG for a chemical exhibiting a noncancer threshold for toxic effects or a
nonlinear carcinogen includes an oral toxicity reference value (RfV) such as an RfD (or MRL),
7 An MRL is an estimate of the daily human exposure to a hazardous substance that is likely to be without
appreciable risk of adverse noncancer health effects over a specified duration of exposure (ATSDR, 2021).
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body weight-based drinking water intake (DWI-BW), and relative source contribution (RSC) as
presented in Equation 1.
MCLG = (°ral RfD) * RSC (Eqn. 1)
V DWI-BW) v M '
Where:
RfD = chronic reference dose—an estimate (with uncertainty spanning perhaps an order
of magnitude) of a daily oral exposure of the human population to a substance that is
likely to be without an appreciable risk of deleterious effects during a lifetime.
DWI-BW = An exposure factor in the form of the 90th percentile body weight-adjusted
drinking water intake for the identified population or life stage, in units of liters of water
consumed per kilogram body weight per day (L/kg/day). The DWI-BW considers both
direct and indirect consumption of drinking water (indirect water consumption
encompasses water added in the preparation of foods or beverages, such as tea or coffee).
Chapter 3 of EPA's Exposure Factors Handbook (EPA, 2019a) provides DWI-BWs for
various populations or life stages within the general population based on publicly
available, peer-reviewed data such as National Health and Nutrition Examination Survey
(NHANES) data.
RSC = relative source contribution—the percentage of the total oral exposure attributed
to drinking water sources (EPA, 2000b), with the remainder of the exposure allocated to
all other routes or sources.
The approach to select the DWI-BW and RSC for MCLG derivation includes a step to identify
sensitive population(s) or life stage(s) (i.e., those that may be more susceptible or sensitive to a
chemical exposure) by considering the available data for the contaminant, including the adverse
health effects in the toxicity study on which the RfD was based (known as the critical effect
within the critical or principal study). Although data gaps can complicate identification of the
most sensitive population (e.g., not all windows or life stages of exposure or health outcomes
may have been assessed in available studies), the critical effect and point of departure8 (POD)
that form the basis for the RfD (or MRL) can provide some information about sensitive
populations because the critical effect is typically observed at the lowest tested dose among the
available data. Evaluation of the critical study, including the exposure window, may identify a
sensitive population or life stage (e.g., pregnant women, formula-fed infants, lactating women).
In such cases, EPA can select the corresponding DWI-BW for that sensitive population or life
stage from the Exposure Factors Handbook (EPA, 2019a) to derive the MCLG. In the absence of
information indicating a sensitive population or life stage, the DWI-BW corresponding to the
general population may be selected for use in MCLG derivation.
To account for potential aggregate risk from exposures and exposure pathways other than oral
ingestion of drinking water, EPA applies an RSC when calculating MCLGs to ensure that an
8 The POD is the dose-response point that marks the starting point for low-dose extrapolation. The POD may be a
no-observed-adverse-effect level (NOAEL)/lowest-observed-adverse-effect level (LOAEL), but ideally is
established from benchmark dose (BMD) modeling of the experimental data, and generally corresponds to a selected
estimated low-level of response (e.g., 1% to 10% incidence for a quantal effect). Depending on the mode of action
and other available data, some form of extrapolation below the POD may be employed for estimating low-dose risk
or the POD may be divided by a series of UFs to arrive at a reference dose (RfD) (EPA, 2012).
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individual's total exposure to a contaminant does not exceed the daily exposure associated with
the toxicity value (threshold level), consistent with EPA (2000b) and long-standing EPA
methodology for establishing drinking water MCLGs and NPDWRs. The RSC represents the
proportion of an individual's total exposure to a contaminant that is attributed to drinking water
ingestion (directly or indirectly in beverages like coffee, tea, or soup, as well as from transfer to
dietary items prepared with drinking water) relative to other exposure pathways. The remainder
of the exposure equal to the RfD (or MRL) is allocated to other potential exposure sources (EPA,
2000b). The purpose of the RSC is to ensure that the level of a contaminant (e.g., MCLG), when
combined with other identified potential sources of exposure for the population of concern, will
not result in exposures that exceed the RfD (or MRL) (EPA, 2000b).
To determine the RSC, EPA follows the Exposure Decision Tree for Defining Proposed RfD (or
POD/UF) Apportionment in EPA's Methodology for Deriving Ambient Water Quality Criteria
for the Protection of Human Health (EPA, 2000b). EPA considers whether there are significant
known or potential uses/sources of the contaminant other than drinking water, the adequacy of
data and strength of evidence available for each relevant exposure medium and pathway, and
whether adequate information on each exposure source is available to quantitatively characterize
the exposure profile. The RSC is developed to reflect the exposure to the general population or a
sensitive population within the general population. When exposure data are available for
multiple sensitive populations or life stages, the most health-protective RSC is selected. In the
absence of adequate data to quantitatively characterize exposure to a contaminant, EPA typically
selects an RSC of 20% (0.2). When scientific data demonstrating that sources and routes of
exposure other than drinking water are not anticipated for a specific pollutant, the RSC can be
raised as high as 80% based on the available data, thereby allocating the remaining 20% to other
potential exposure sources (EPA, 2000b).
EPA's protocol for MCLG development for the mixture of HFPO-DA, PFBS, PFNA, and
PFHxS follows existing agency guidance, policies, and procedures related to the three key inputs
described above (i.e., RfD/MRL, DWI-BW, and RSC) and longstanding agency mixtures
guidance (EPA, 1986, 2000a) to address dose additive effects. First, EPA derives a health-based
water concentration (HBWC), calculated using the MCLG equation above (Equation 1), for each
of the four PFAS (see Sections IIA-D). Peer reviewed, publicly available assessments for
HFPO-DA (EPA, 2021c), PFBS (EPA, 2021d), PFNA (ATSDR, 2021), and PFHxS (ATSDR,
2021) provide the oral toxicity values (i.e., RfD or MRL) used to calculate the HBWCs for these
four PFAS. The DWI-BW for each of the four PFAS is selected from EPA's Exposure Factors
Handbook (EPA, 2019a), taking into account the relevant sensitive population(s) or life stage(s).
RSCs are determined based on a literature review of potential exposure sources of the four
PFAS, and analysis using the Exposure Decision Tree approach (EPA, 2000b). Finally, to
account for dose additive noncancer effects, the HBWCs for HFPO-DA, PFBS, PFNA, and
PFHxS are used in a HI approach, a commonly used component-based mixture risk assessment
method (EPA, 2022c).
Following EPA's data-driven approach for component-based mixtures assessment based on dose
additivity (i.e., see Figure 4-1 in EPA, 2022c), the agency selected the HI approach for MCLG
development because HBWCs for HFPO-DA, PFBS, PFNA, and PFHxS are available or can be
calculated. Although a single PFAS may occur in drinking water at concentrations below where
EPA might establish an individual MCLG, PFAS tend to co-occur (see Section I.B). Hence,
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deriving a MCLG based on the concentration of an individual PFAS without consideration of the
dose additive effects that would occur from other PFAS that may be present in water may not
result in a sufficiently protective MCLG with an adequate margin of safety. The HI approach is
health-protective for development of a mixture MCLG because the HBWCs are based on the
most sensitive known adverse health outcome for each mixture component. Thus, the overall HI
ensures that the MCLG protects against noncancer health risk associated with exposure to a
mixture of PFHxS, HFPO-DA, PFNA, and PFBS.
The HI is based on an assumption of dose addition (DA) among the mixture components (EPA,
2000a; Svendsgaard and Hertzberg, 1994). In the HI approach, an HQ is calculated as the ratio of
human exposure (E) to a health-based RfV for each mixture component chemical (i) (EPA,
1986). The HI involves the use of RfVs for each PFAS mixture component (in this case, PFHxS,
HFPO-DA, PFNA, and PFBS), which have been selected based on sensitive health outcomes and
which are expected to be protective of all other adverse health effects observed after exposure to
the individual PFAS. This approach, which protects against all adverse effects and not just a
single adverse outcome/effect, is a conservative risk indicator and appropriate for MCLG
development. The HI is dimensionless, so in the HI formula, E and the RfV must be in the same
units (Equation 2). For example, if E is the oral intake rate (milligrams per kilogram per day
(mg/kg/day)), then the RfV could be the RfD or MRL, which have the same units. Alternatively,
the exposure metric can be a media-specific metric such as a measured water concentration (e.g.,
nanograms per liter or ng/L) and the RfV can be an HBWC (e.g., ng/L). The component
chemical HQs are then summed across the mixture to yield the HI (Equation 2). A mixture HI
exceeding 1.0 indicates potential risk for a given environmental medium or site. The HI provides
an indication of: (1) concern for the overall mixture and (2) potential driver PFAS (i.e., those
PFAS with high(er) HQs). For a detailed discussion of PFAS dose additivity and the HI
approach, see the PFAS Mixtures Framework (EPA, 2022c).
HI = Hazard Index
HQ, = Hazard Quotient for chemical i
Ei = Exposure, i.e., dose (mg/kg/day) or occurrence concentration, such as in drinking
water (in milligrams per liter or mg/L), for chemical i
RfVi = Reference value (e.g., oral RfD or MRL) (mg/kg/day), or corresponding HBWC;
e.g., such as a MCLG for chemical i (in mg/L)
n n £
(Eqn. 2)
Where:
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2.0 Health-Based Water Concentrations
2.1 HFPO-DA
HFPO dimer acid and its ammonium salt are shorter-chain PFAS that were intended to be a
replacement for the longer-chained PFOA. Together, they are referred to as "GenX chemicals"
because they are associated with the GenX processing aid technology to make fluoropolymers
without using PFOA (EPA, 2021c). In water, both HFPO dimer acid and its ammonium salt
dissociate to form the HFPO dimer acid anion (HFPO-) as a common analyte.
The HBWC that the agency is using for the HI MCLG was derived from the agency's 2021
human health toxicity assessment, specifically the chronic RfD of 3E-06 mg/kg/day based on
liver effects observed following oral exposure of mice to GenX Chemicals (EPA, 2021c).
Summaries of key information from the toxicity assessment and health advisory document (EPA,
2022a; i.e., information about the RfD, DWI-BW, and RSC that were used to derive the lifetime
noncancer HA value for HFPO-DA) are presented in the following sections. Based on the
toxicity assessment, the HBWC value of 10 ng/L for HFPO-DA is used as a component of the HI
MCLG for the mixture of HFPO-DA, PFBS, PFNA, and PFHxS (see Section 3 .0).
2.1.1 Toxicity
EPA's HBWC for HFPO-DA is derived from a chronic RfD that is based on liver effects
observed following oral exposure of mice to HFPO-DA (EPA, 2021c, 2022a).
Oral toxicity studies in rodents exposed to HFPO-DA report a range of toxic effects. Repeated-
dose oral exposure of rats and mice resulted in liver toxicity (e.g., increased relative liver weight,
hepatocellular hypertrophy, apoptosis, and single-cell/focal necrosis), kidney toxicity (e.g.,
increased relative kidney weight), immune system effects (e.g., antibody suppression),
hematological effects (e.g., decreased red blood cell count, hemoglobin, and hematocrit),
reproductive/developmental effects (e.g., increased number of early deliveries, placental lesions,
changes in maternal gestational weight gain, and delays in genital development in offspring), and
cancer (e.g., liver and pancreatic tumors) (EPA, 2021c).
The most sensitive noncancer effects observed among the available data were the adverse effects
on liver, which were observed in both male and female mice and rats across a range of exposure
durations and dose levels, including the lowest tested dose levels and shortest exposure durations
(EPA, 2021c). Noncancer liver effects formed the basis for the chronic RfD of 3E-06 mg/kg/day,
which EPA used to derive the lifetime HA value of 10 ng/L for HFPO-DA (EPA, 2022a). To
develop the chronic RfD for HFPO-DA, EPA derived a human equivalent dose (HED) of
0.01 mg/kg/day from a no-observed-adverse-effect level (NOAEL) of 0.1 mg/kg/day for liver
effects in the identified critical study (an oral reproductive/developmental toxicity study in mice;
Dupont 18405-1037, 2010). EPA then applied a composite uncertainty factor (UF) of 3,000 (i.e.,
10X for intraspecies variability (UFh), 3X for interspecies differences (UFa), 10X for
extrapolation from a subchronic to a chronic dosing duration (UFs), and 10X for database
deficiencies (UFd)) to yield the chronic RfD (EPA, 2021c).
There is suggestive evidence of carcinogenic potential of oral exposure to HFPO-DA in humans,
but the available data are insufficient to derive a cancer risk concentration in water for HFPO-
DA (EPA, 2021c, 2022a).
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2.1.2 Exposure Factor
To select an appropriate DWI-BW for use in derivation of the noncancer HBWC for HFPO-DA,
EPA considered the HFPO-DA exposure interval used in the oral reproductive/developmental
toxicity study in mice that was the basis for chronic RfD derivation (the critical study). In this
study, parental female mice were dosed from pre-mating through lactation, corresponding to
three potentially sensitive human adult life stages that may represent critical windows of
exposure for HFPO-DA: women of childbearing age, pregnant women, and lactating women
(Table 3-63 in EPA, 2019a). Of these three, the DWI-BW for lactating women (0.0469 L/kg/day)
is anticipated to be protective of the other two sensitive life stages. Therefore, EPA used the
DWI-BW for lactating women to calculate the noncancer lifetime HA value for HFPO-DA
(EPA, 2022a).
2.1.3 Relative Source Contribution
The HBWC for HFPO-DA was calculated using an RSC of 0.20, meaning that 20% of the
exposure—equal to the RfD—is allocated to drinking water, and the remaining 80% is attributed
to all other potential exposure sources (EPA, 2022a). Selection of this RSC was based on EPA's
determination that the available exposure data for HFPO-DA did not enable a quantitative
characterization of relative HFPO-DA exposure sources and routes. In such cases, an RSC of
0.20 is typically used (EPA, 2000b).
2.1.4 Derivation of HFPO-DA HBWC
The HBWC for HFPO-DA and is calculated as follows and summarized in Table 1:
/ RfD \
GenX Chemicals HBWC = —-—— * RSC
\DWI-B W /
DWI-BW
= 10 — or parts per trillion (ppt)
Li
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Table 1. HFPO-DA HBWC - Input Parameters and Value
Parameter Value Units Source
Chronic 3E-06 mg/kg/day Final RfD based on critical liver effects (constellation of liver
oral RfD lesions as defined by the National Toxicology Program
(NTP) Pathology Working Group) in parental female mice
exposed to HFPO dimer acid ammonium salt by gavage from
pre-mating through lactation (53-64 days) (EPA, 2021c;
Dupont 18405-1037, 2010).
DWI-BW 0.0469 L/kg/day 90th percentile two-day average, consumer-only estimate of
combined direct and indirect community water ingestion for
lactating women (age 13 to < 50 years) based on 2005-2010
NHANES data (EPA, 2019a).
RSC 0.2 N/A Based on a review of the available scientific literature on
HFPO-DA potential exposure routes and sources (EPA,
2021c).
HFPO-DA HBWC = 0.00001 mg/L or 10 ppt
2.2 PFBS
PFBS and its related compound K+PFBS are shorter-chain PFAS that were developed as "safer"
replacements for the longer-chained PFOS. In water, K+PFBS dissociates to the deprotonated
anionic form of PFBS (PFBS-) and the K+ cation at environmental pH levels (pH 4-9). These
three PFBS chemical forms are referred to collectively as PFBS.
The HBWC that the agency is using for the HI MCLG was derived from the agency's 2021
human health toxicity assessment, specifically the chronic RfD of 3E-04 mg/kg/day based on
thyroid effects observed seen in newborn mice born to mothers that had been orally exposed to
PFBS throughout gestation (EPA, 202Id). Summaries of key information from the toxicity
assessment and HA document (i.e., information about the RfD, DWI-BW, and RSC that were
used to derive the lifetime noncancer HA value for PFBS) are presented in the following
sections. Based on the toxicity assessment, and consistent with the HA analysis, the HBWC of
2,000 ng/L for PFBS is used as a component of the Hazard Index MCLG for the mixture of
HFPO-DA, PFBS, PFNA, and PFHxS (see Section 3.0).
2.2.1 Toxicity
EPA's HBWC for PFBS was derived using a chronic oral RfD based on thyroid effects seen in
an oral toxicity study in mice (EPA, 202Id, 2022b).
EPA's final toxicity assessment for PFBS (EPA, 202Id) considered all publicly available human,
animal, and mechanistic studies of PFBS exposure and effects. The assessment identified
associations between PFBS exposure and thyroid, developmental, and kidney effects based on
studies in animals. The limited evidence for thyroid or kidney effects in human studies was
equivocal, and no studies evaluating developmental effects of PFBS in humans were available.
Human and animal studies evaluated other health effects following PFBS exposure including
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effects on the reproductive system, liver, and lipid and lipoprotein homeostasis, but the evidence
did not support clear associations between exposure and effect (EPA, 202Id).
The most sensitive noncancer effect observed was an adverse effect on the thyroid (i.e.,
decreased serum total thyroxine) seen in newborn mice (postnatal day (PND) 1) born to mothers
that had been orally exposed to K+PFBS throughout gestation (Feng et al., 2017; EPA, 2021d).
This critical effect was the basis for the chronic RfD of 3E-04 mg/kg/day which EPA used to
derive the noncancer lifetime HA value for PFBS (EPA, 202Id, 2022b). To develop the chronic
RfD for PFBS,9 EPA derived an HED of 0.095 mg/kg/day from benchmark dose (BMD)
modeling of the critical effect in mice. EPA then applied a composite UF of 300 (i.e., 10X for
UFh, 3X for UFa, and 10X for UFd) to yield the chronic RfD (EPA, 202Id).
There were no human or animal studies identified that evaluated the potential carcinogenicity of
PFBS (EPA, 202Id, 2022b).
2.2.2 Exposure Factor
To select an appropriate DWI-BW for use in deriving the HBWC, EPA considered the PFBS
exposure interval used in the developmental toxicity study in mice that was the basis for chronic
RfD derivation. In this study, pregnant mice were exposed throughout gestation, which is
relevant to two human adult life stages: women of child-bearing age who may be or become
pregnant, and pregnant women and their developing embryo or fetus (Table 3-63 in EPA,
2019a). Of these two, EPA selected the DWI-BW for women of child-bearing age
(0.0354 L/kg/day) to derive the noncancer lifetime HA for PFBS because it is higher and
therefore more health-protective (EPA, 2022b).
2.2.3 Relative Source Contribution
The HBWC for PFBS was calculated using an RSC of 0.20, meaning that 20% of the exposure—
equal to the RfD—is allocated to drinking water, and the remaining 80% is attributed to all other
potential exposure sources (EPA, 2022b). This was based on EPA's determination that the
available data on PFBS exposure routes and sources did not enable a quantitative characterization
of PFBS exposure. In such cases, an RSC of 0.20 is typically used (EPA, 2000b).
2.2.4 Derivation of PFBS HB WC
The HBWC for PFBS and is calculated as follows and summarized in Table 2:
9 Data for K+PFBS were used to derive the chronic RfD for the free acid (PFBS), resulting in the same value (3E-
04 mg/kg/day), after adjusting for differences in molecular weight (MW) between K+ PFBS (338.19) and PFBS
(300.10) (EPA, 202Id).
/ RfD \
PFBS HBWC = * RSC
VDWI-BW /
DWI-BW
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mg / mg\
= 0.0017 —— (rounded to 0.002 ——J
Lj ^ Lj '
= 2,000 — or ppt
Li
Table 2. PFBS HBWC - Input Parameters and Value
Parameter
Value
Units
Source
Chronic
RfD
3E-04
mg/kg/day Final RfD based on critical effect of decreased serum total
thyroxine in newborn (PND 1) mice after gestational
exposure to the mother (EPA, 2021d; Feng et al., 2017).
DWI-BW
0.0354
L/kg/day
90th percentile two-day average, consumer-only estimate of
combined direct and indirect community water ingestion for
women of childbearing age (13 to < 50 years) based on
2005-2010 NHANES data (EPA, 2019a).
RSC
0.2
N/A
Based on a review of the available scientific literature on
PFBS potential exposure routes and sources (EPA, 2021d).
PFBS HBWC = 0.002 mg/L or 2,000 ppt
2.3 PFNA
PFNA has been used as a processing aid in the production of fluoropolymers, primarily
polyvinylidene fluoride (PVDF), which is a plastic designed to be temperature resistant and
chemically nonreactive (EPA, 2020; NJDWQI, 2017; Prevedouros et al., 2006). PFNA has been
used since the 1950s in a wide variety of industrial and consumer products (see RSC Section
below). It has also been used in aqueous film-forming foam (AFFF) for fire suppression (EPA,
2020; Laitinen et al., 2014).
PFAS have been measured in human blood samples taken as part of the NHANES. PFNA was
measured in serum samples collected in 2013-2014 from more than 2,000 survey participants,
with a geometric mean concentration of 0.675 micrograms per liter (|ig/L) and 95th percentile
concentration of 2.00 |ig/L (EPA, 202le).
ATSDR has published a toxicological profile for a group of PFAS including PFNA and has
developed an intermediate-duration oral MRL for PFNA (ATSDR, 2021)10. EPA's derived
HBWC for PFNA (described below) is based on the ATSDR MRL (ATSDR, 2021), a DWI-BW
(selected by EPA) that corresponds to this MRL, and an RSC selected by EPA. There is no
published EPA human health toxicity assessment for PFNA; however, EPA's Integrated Risk
Information System (IRIS) program is developing a human health toxicity assessment for PFNA,
10 ATSDR is currently updating their assessment for PFNA, and their perfluoroalkyls assessment is "in
development" (https://www.atsdr.cdc.gov/toxprofiledocs/index.html').
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which is expected to undergo public comment and external peer review in FY2023 (EPA, 202le,
2022c). EPA's IRIS assessment will use systematic review methods to evaluate the
epidemiological and toxicological literature for PFNA, including consideration of relevant
mechanistic evidence (EPA, 202le).
2.3.1 Toxicity
The HBWC for PFNA is based on an ATSDR intermediate-duration oral MRL that was based on
developmental effects seen in mice after oral PFNA exposure (ATSDR, 2021).
Studies of oral PFNA exposure in rodents have reported adverse effects on the liver,
development, and reproductive and immune systems (ATSDR, 2021). The most sensitive
noncancer effects and basis for the ATSDR intermediate-duration oral MRL (and thus EPA's
HBWC) were decreased body weight gain and impaired development (i.e., delayed eye opening,
preputial separation, and vaginal opening) in mice born to mothers that were gavaged with
PFNA from gestational days (GDs) 1-17 (with presumed continued indirect exposure of
offspring via lactation), and monitoring until PND 287 (ATSDR, 2021). The study reporting
these effects (Das et al., 2015) was selected by ATSDR as the principal study for MRL
derivation. To derive the MRL, an HED of 0.001 mg/kg/day was calculated from the NOAEL of
1 mg/kg/day identified in the study. Then, ATSDR applied a total UF of 30 (i.e., 10X for UFh
and 3X for UFa) and a modifying factor (MF) of 10X for database deficiencies to account for the
small number/limited scope of studies examining PFNA toxicity following intermediate-duration
exposure. The resulting intermediate-duration oral MRL was 3E-06 mg/kg/day (ATSDR, 2021).
EPA did not apply an additional UF to adjust for subchronic-to-chronic duration (i.e., UFs) to
calculate the HBWC because the critical effects were observed during a developmental life
stage11 (EPA, 2002). Toxicological assessments based on animal studies for PFNA from other
sources (e.g., states) are in a similar range as this value, providing additional support (e.g.,
4.3 x 10"6 mg/kg/day to 1.2 x 10"5 mg/kg/day; see Addendum A in EPA, 2021e).
The carcinogenic potential of PFNA has been examined in three epidemiological studies. No
consistent associations between serum PFNA levels and breast cancer or prostate cancer were
found (ATSDR, 2021).
2.3.2 Exposure Factor
Based on the life stages of exposure in the principal study from which the intermediate-duration
MRL was derived (i.e., directly to maternal animals during gestation, and indirectly to offspring
during gestation and lactation), EPA identified three potentially sensitive life stages that may
represent critical windows of exposure for PFNA: women of childbearing age (13 to < 50 years),
pregnant women, and lactating women (Table 3-63 in EPA, 2019a). The DWI-BW for lactating
women (0.0469 L/kg/day; 90th percentile direct and indirect consumption of community water,
consumer-only two-day average) was selected to calculate the HBWC for PFNA because it is the
11 As stated in EPA (2002), "... This is because it is assumed that most endpoints of developmental toxicity can be
caused by a single exposure. If, however, developmental effects are more sensitive than those seen after longer-term
exposures, then even the chronic RfD/RfC should be based on such effects to reduce the risk of potential greater
sensitivity in children. Because the standard studies currently conducted for developmental toxicity involve repeated
exposures, data are not often available on which endpoints may be induced by acute, subacute, subchronic, or
chronic dosing regimens and, therefore, on which should be used in setting various duration reference values."
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highest of the three DWI-BWs and is anticipated to be protective of the other two sensitive life
stages.
2.3.3 Relative Source Contribution
EPA calculated the HBWC for PFNA using an RSC of 0.20, meaning that 20% of the
exposure—equal to the MRL—is allocated to drinking water, and the remaining 80% is
attributed to all other potential exposure sources. This was based on EPA's determination that
the available data on PFNA exposure routes and sources did not permit quantitative
characterization of PFNA exposure. In such cases, an RSC of 0.20 is typically used (EPA,
2000b). See Appendix A for complete details on the RSC determination for PFNA.
2.3.4 Derivation of PFNA HB WC
The HBWC for PFNA is calculated as follows and summarized in Table 3:
MRL
PFNA HBWC = (¦
/0.000003
* RSC
VDWI-BW
mg \
1.0.2
\ 0.0469, ,L, /
\ kg/day /
mg / mg\
= 0.000014 —— (rounded to 0.00001 ——J
Lj ^ Lj '
= 0.01 ^
Li
ng
= 10 — or ppt
Li
Table 3. PFNA HBWC - Input Parameters and Value
Parameter
Value
Units
Source
Intermediate-
Duration Oral
MRL
3E-06a
mg/kg/day
Based on decreased body weight gain and delayed eye
opening, preputial separation, and vaginal opening in
mouse offspring after gestational and presumed
lactational exposure (ATSDR, 2021; Das et al., 2015).
DWI-BW
0.0469
L/kg/day
90th percentile two-day average, consumer-only
estimate of combined direct and indirect community
water ingestion for lactating women (13 to < 50 years)
based on 2005-2010 NHANES data (EPA, 2019a).
RSC
0.2
N/A
Based on a review of the current scientific literature
summarized in this document (see Appendix A).
PFNA HBWC = 0.00001 mg/L or 10 ppt
Note:
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aNote that ATSDR MRLs and EPA RfDs are not necessarily equivalent (e.g., intermediate-duration MRL vs. chronic RfD; EPA
and ATSDR may apply different uncertainty/modifying factors) and are developed for different purposes. In this case, EPA did
not apply an additional LTF s to calculate the HBWC for PFNA because the critical effect is identified in a developmental
population (EPA, 2002).
2.4 PFHxS
PFHxS has been used in laboratory applications and as a raw material or a precursor for the
manufacture of PFAS/perfluoroalkyl sulfonate-based products, though production of PFHxS in
the United States was phased out by its major manufacturer in 2002 (Backe et al., 2013; Buck et
al., 2011; OECD, 2011 and Sigma-Aldrich, 2014 as cited in NCBI, 2022). PFHxS has also been
used in firefighting foam and carpet treatment solutions, and it has been used as a stain and water
repellant (Garcia and Harbison, 2015 as cited in NCBI, 2022).
PFAS have been measured in human blood samples taken as part of the NHANES. PFHxS was
measured in serum samples collected in 2013-2014 from more than 2,000 survey participants,
with a geometric mean concentration of 1.35 |ig/L and 95th percentile concentration of 5.60 |ig/L
(EPA, 202le).
ATSDR has published a toxicological profile for a group of PFAS including PFHxS and has
calculated an intermediate-duration oral MRL for PFHxS (ATSDR, 2021)12. EPA's derived
HBWC for PFHxS described below is based on the ATSDR MRL (ATSDR, 2021), a DWI-BW
(selected by EPA) that corresponds to the MRL, and an RSC selected by EPA. There is no
published EPA human health toxicity assessment for PFHxS; however, EPA's IRIS program is
developing a human health toxicity assessment for PFHxS, which is expected to undergo public
comment and external peer review in FY2023 (EPA, 2022c). EPA's IRIS assessment will use
systematic review methods to evaluate the epidemiological and toxicological literature for
PFHxS, including consideration of relevant mechanistic evidence (EPA, 202le).
2.4.1 Toxicity
The HBWC for PFHxS is derived using an ATSDR intermediate-duration oral MRL based on
thyroid effects seen in male rats after oral PFHxS exposure (ATSDR, 2021). Toxicity studies of
oral PFHxS exposure to animals also have reported health effects on the liver, thyroid, and
development (ATSDR, 2021). The most sensitive noncancer effect observed was thyroid
follicular epithelial hypertrophy/hyperplasia in parental male rats that had been exposed for 42-
44 days, identified in the principal developmental toxicity study selected by ATSDR (NOAEL of
1 mg/kg/day for this effect) (Butenhoff et al., 2009; ATSDR, 2021). This critical effect was the
basis for the ATSDR intermediate-duration oral MRL which EPA used to derive the HBWC for
PFHxS. An HED of 0.0047 mg/kg/day was calculated from the NOAEL of 1 mg/kg/day
identified in the principal study. ATSDR applied a total UF of 30 (i.e., 10X for UFh and 3X for
UFa) and a MF of 10X for database deficiencies to yield an intermediate-duration oral MRL of
2E-05 mg/kg/day (ATSDR, 2021). To calculate the HBWC, EPA applied an additional UF of 10
to adjust for subchronic-to-chronic duration (i.e., UFs), per agency guidance (EPA, 2002),
because the effect is not in a developmental population (i.e., thyroid follicular epithelial
hypertrophy/hyperplasia in parental male rats). The resulting adjusted chronic reference value is
2E-06 mg/kg/day. Toxicological assessments based on animal studies for PFHxS from other
12 ATSDR is currently updating their assessment for PFHxS, and their perfluoroalkyls assessment is "in
development" (https://www.atsdr.cdc.gov/toxprofiledocs/index.html').
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sources (e.g., states) are in a similar range as this value, providing additional support (e.g.,
3.8 x 10"6 to 9.7 x 10"6 mg/kg/day; see Addendum A in EPA, 202le).
The carcinogenic potential of PFHxS has been examined in four epidemiological studies
(ATSDR, 2021). Bonefeld-Jorgensen et al. (2014) reported a significant negative correlation
between serum PFHxS levels (mean concentration 1.2 ng/mL) and breast cancer risk among
Danish women. However, a study in Greenland found a significant, positive association between
high serum levels of PFHxS and breast cancer risk (Wiels0e et al., 2017). The median serum
PFHxS concentration among cases in that study was 2.52 ng/mL and serum levels ranged from
0.19 ng/mL to 23.40 ng/mL (Wiels0e et al., 2017). Hardell et al. (2014) found a statistically
significant interaction between above-median PFHxS concentrations and increased risk for
prostate cancer among men with genetics as a risk factor (first-degree relative). Prostate-specific
antigen (PSA) levels were not associated with serum PFHxS levels (mean concentration
3.38 ng/mL) in men 20-49 or 50-69 years of age (Ducatman et al., 2015). EPA has not yet
completed an evaluation and classification of the carcinogenicity of PFHxS, and thus, the
HBWC and MCLG are based on noncancer effects.
2.4.2 Exposure Factor
No sensitive population or life stage was identified for DWI-BW selection for PFHxS because
the critical effect on which the ATSDR MRL was based (thyroid alterations) was observed in
adult male rats. Since this exposure life stage does not correspond to a sensitive population or
life stage, a DWI-BW for adults within the general population (0.034 L/kg/day; 90th percentile
direct and indirect consumption of community water, consumer-only two-day average, adults 21
years and older) was selected for HBWC derivation (EPA, 2019a).
2.4.3 Relative Source Contribution
EPA calculated the HBWC for PFHxS using an RSC of 0.20, meaning that 20% of the
exposure—equal to the chronic reference value—is allocated to drinking water, and the
remaining 80% is attributed to all other potential exposure sources. This was based on EPA's
determination that the available data on PFHxS exposure routes and sources did not permit
quantitative characterization of PFHxS exposure. In such cases, an RSC of 0.20 is typically used
(EPA, 2000b). See Appendix B for complete details on the RSC determination for PFHxS.
2.4.4 Derivation of PFHxS HB WC
The HBWC for PFHxS is calculated as follows and summarized in Table 4:
PFHxS HBWC = (
Chronic reference value
DWI-BW
= 0.0000092 ^ (rounded to 0.000009
Lj ^ Lj '
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= 0.009 -f
J_j
ng
= 9 — or ppt
Li
Table 4. PFHxS HBWC - Input Parameters and Value
Parameter
Value
Units
Source
Chronic
reference value
2E-06a
mg/kg/day
Based on thyroid follicular epithelial
hypertrophy/hyperplasia in parental male rats
(exposed 42-44 days) (ATSDR, 2021;
Butenhoff et al., 2009).
DWI-BW
0.034
L/kg/day
90th percentile two-day average, consumer-
only estimate of combined direct and indirect
community water ingestion for adults
21 years and older based on 2005-2010
NHANES data (EPA, 2019a).
RSC
0.2
N/A
Based on a review of the current scientific
literature summarized in this document (see
Appendix B).
PFHxS HBWC = 0.000009 mg/L or 9 ppt
Note:
aNote that MRLs and RfDs are not necessarily equivalent (e.g., intermediate-duration MRL vs. chronic RfD, EPA and ATSDR
may apply different uncertainty /modifying factors) and are developed for different purposes. In this case, EPA applied an
additional LTF of 10 to account for subchronic-to-chronic duration (i.e., LTFs) yielding a chronic reference value of 2E-
06 mg/kg/day, which was used to calculate the HBWC for PFHxS (EPA, 2002).
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To account for dose additive noncancer effects associated with HFPO-DA, PFBS, PFHxS, and
PFNA, EPA is proposing a MCLG for the mixture of these four PFAS based on the HI approach
(EPA, 2022c). As described in Section I.D., a mixture HI can be calculated when HBWCs (e.g.,
HAs) for a set of PFAS are available or can be calculated. HQs are calculated by dividing the
measured component PFAS concentration in water (e.g., expressed as ng/L) by the relevant
HBWC (e.g., expressed as ng/L), as shown in the equation below. Component HQs are then
summed across the PFAS mixture to yield the PFAS mixture HI MCLG. A PFAS mixture HI
MCLG greater than 1.0 indicates an exceedance of the health protective level and indicates
potential human health risk for noncancer effects from the PFAS mixture in water. For more
details, please see EPA (2022c). The proposed mixture HI MCLG for HFPO-DA, PFBS, PFNA,
and PFHxS is as follows:
HI MCLG ( ^^-^waterl ^ _|_ f [PFBSwater] \ _|_ f [PFNAwater! \ _|_ ( [PFHxSwafer] A j q
\[GenXHBwc]<' \[PFBSHbwc\) ylPFNAMBWc]-^ \[PFHxShbwcV
HI MCLG water l\ _|_ f[PFBSwa(er]\ f[PFNAwa(er]\ ^ /[PFHxSwafer] A j q
~ V [10 ng/L] J V[2000 ng/L]) V [10 ng/L] J V [9ng/L] J ~
Where
[PFASwater] = the measured component PFAS concentration in water and
[PFAShbwc] = the HBWC of a component PFAS.
Although current weight of evidence suggests that PFAS vary in their precise structure and
function, exposure to different PFAS can result in similar health effects; as a result, PFAS
exposures are likely to result in dose-additive effects and therefore the assumption of dose-
additivity is reasonable (ATSDR, 2021; EPA, 2022a). While individual PFAS can pose a
potential risk to human health if the exposure level exceeds the chemical-specific toxicity value
(RfD or MRL) (i.e., individual PFAS HQ > 1.0), mixtures of PFAS can result in dose-additive
health effects when lower individual concentrations of PFAS are present in that mixture. For
example, if the individual HQs for PFHxS, HFPO-DA, PFNA, and PFBS were each 0.9 that
would indicate that the measured concentration of each PFAS in drinking water is below the
level of appreciable risk (recall that an RfV, such as an oral RfD, represents an estimate at which
no appreciable risk of deleterious effects exists). However, the overall HI for that mixture would
be 3.6 (i.e., sum of four HQs of 0.9). A HI of 3.6 means that the total measured concentration of
PFAS is 3.6 times the level associated with potential health risks. Thus, setting a MCLG based
on the concentration of an individual PFAS without considering the potential dose-additive
effects from other PFAS in a mixture would likely not provide a sufficiently protective MCLG
with an adequate margin of safety. In order to account for dose additive noncancer effects
associated with co-occurring PFAS, to protect against health impacts from likely multi-chemical
exposures of PFHxS, HFPO-DA, PFNA, and PFBS, the agency is proposing use of the HI
approach, a commonly used component-based mixture risk assessment method, for the MCLG
for these four PFAS. Consistent with the statutory requirement under 1412(b)(4)(A) of SDWA,
establishing the MCLG for PFHxS, HFPO-DA, PFNA, and PFBS at a HI = 1.0 ensures that the
MCLG is set at a level at which there are no known or anticipated adverse effect on the health of
persons and which ensures an adequate margin of safety.
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EPA (U.S. Environmental Protection Agency). 2023c. Public Comment Draft - Toxicity
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PFOS in Drinking Water. EPA-822-P-23-008. EPA, Office of Water, Washington, DC.
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NJDWQI (New Jersey Drinking Water Quality Institute). 2017. Public review draft: Health-
based maximum contaminant level support document: Perfluorooctane sulfonate (PFOS)
(CAS #: 1763-23-1; Chemical Formula: C8HF1703S). NJDWQI Health Effects
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Docum entati on - Supportin g-PF OS-13 -ppt.pdf.
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APPENDIX A. PFNA: Summary of Occurrence in Water and Detailed
Relative Source Contribution.
Occurrence in Water
The use and production of PFNA could result in its release to the aquatic environment through
various waste streams (NCBI, 2022). PFNA has an estimated water solubility of 62.5 |ig/L
(6.25 x 10"2 mg/L) at 25 °C and when released to surface water, it is expected to adsorb to
suspended solids and sediment (NCBI, 2022). Volatilization from water surfaces is not expected
to be an important fate process for PFNA (NCBI, 2022).
Drinking Water
Based on results from EPA's UCMR 3 monitoring, PFNA has been detected in 0.28% of
drinking water systems in the United States, with mean and median detected concentrations of
36 ng/L and 32 ng/L, respectively. The UCMR 3 maximum concentration detected in drinking
water systems was 56 ng/L (EPA, 2017). Drinking water samples collected from public drinking
water systems that are impacted by wastewater treatment effluent often contain higher
concentrations of perfluoroalkyls than samples collected from systems that are not impacted by
wastewater treatment effluent (Schultz et al., 2006a,b, as cited in ATSDR, 2021). For example,
PFNA was detected in all samples collected from a public drinking water system in Los Angeles
that was highly impacted by wastewater treatment effluent; the mean concentrations of PFNA in
influent and effluent samples were 5.5 ng/L and 3.5 ng/L, respectively (Quinones and Snyder,
2009 as cited in ATSDR, 2021). In comparison, no perfluoroalkyl chemicals were detected in
influent or finished drinking water in samples collected from a public drinking water system in
Aurora, Colorado that was not highly impacted by wastewater treatment effluent (Quinones and
Snyder, 2009 as cited in ATSDR, 2021). PFNA was detected in 30% of public drinking water
systems tested in New Jersey (Post et al., 2013 as cited in ATSDR, 2021).
A Standard of Quality for PFAS in bottled water is 0.005 |ig/L (5 ng/L) for one PFAS (e.g.,
PFNA) and 0.0010 |ig/L (10 ng/L) for more than one PFAS (IBWA, 2022).
For more information about PFNA occurrence in drinking water, please see EPA (2022f).
Groundwater
PFNA was detected at a concentration of 25.7 ng/L (0.0257 |ig/L) in one of 19 well water
samples collected from farms near Decatur, Alabama that have historically applied
fluorochemical industry-impacted biosolids to fields (Lindstrom et al., 2011 as cited in NCBI,
2022). PFNA was also detected in groundwater samples collected in 2010 from the Highland
Creek watershed in Canada at concentrations ranging from 0.071 ng/L to 0.54 ng/L
(0.000071 |ig/L to 0.00054 |ig/L) (Meyer et al., 2011, as cited in NCBI, 2022). In this study, the
authors reported that none of the sampling sites receive water that is impacted by known PFAS
sources (Meyer et al., 2011). Median and maximum groundwater concentrations of 105 ng/L
(0.105 |ig/L) and 3,000 ng/L (3 |ig/L), respectively, were detected at 10 U.S. military
installations (Anderson et al., 2016 as cited in ATSDR, 2021).
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Surface Water
In a 2016 study, PFNA was detected in each of 37 surface water sampling sites in the northeast
United States, with a maximum concentration of 14 ng/L (0.014 |ig/L) measured at Mill Cove,
Rhode Island (Zhang et al., 2016 as cited in ATSDR, 2021). Concentrations of PFNA in 11 lake
water samples and 14 surface water samples collected in Albany, New York ranged from not
detected to 3.51 ng/L (0.00351 |ig/L) and from < 0.25 ng/L to 5.90 ng/L ( < 0.00025 |ig/L to
0.00590 |ig/L), respectively (Kim and Kannan, 2007 as cited in ATSDR, 2021). Concentrations
of PFNA in 12 remote and urban Minnesota surface water samples, including samples collected
from Lake Michigan, ranged from < 0.3 ng/L to 3.1 ng/L (< 0.0003 |ig/L to 0.0031 |ig/L)
(Simcik and Dorweiler, 2005 as cited in ATSDR, 2021). PFNA was detected in 38% of 8 surface
water samples from U.S. streams in the Great Lakes basin collected during 1994 to 2000 at
concentrations between 0.03 ng/L and 0.4 ng/L (0.00003 |ig/L and 0.0004 |ig/L) (Klecka et al.,
2010 as cited in NCBI, 2022).
PFNA was detected in 6 locations in the Delaware River at concentrations ranging from
1.65 ng/L to 976 ng/L (0.00165 |ig/L to 0.976 |ig/L) in 2007 to 2009 (DRBC, 2013 as cited in
ATSDR, 2021). PFNA was detected at concentrations of 2.24 ng/L to 194 ng/L (0.00224 |ig/L to
0.194 |ig/L) in 11 samples with the highest total PFAS levels out of 100 samples collected from
80 sites in the Cape Fear River Basin, North Carolina in 2006 (Nakayama et al., 2007 as cited in
ATSDR, 2021 and NCBI, 2022). Mean and median PFNA concentrations were 33.6 ng/L and
5.70 ng/L (0.0336 |ig/L and 0.00570 |ig/L), respectively, with PFNA not detected in 10.1% of
the samples (Nakayama et al. 2007 as cited in ATSDR, 2021). PFNA was detected at
concentrations between 12.4 ng/L and 286 ng/L (0.0124 |ig/L to 0.286 |ig/L) in 9 of 32 surface
water samples collected from ponds and streams near farms near Decatur, Alabama that have
historically applied fluorochemical industry-impacted biosolids to fields (Lindstrom et al., 2011
as cited in NCBI, 2022). Median and maximum surface water concentrations of 96 ng/L
(0.096 |ig/L) and 10,000 ng/L (10 |ig/L), respectively, were detected at 10 U.S. military
installations (Anderson et al., 2016 as cited in ATSDR, 2021).
Concentrations of PFNA in creek and river samples measured throughout Canada ranged
from < 125 pg/L to 3,000 pg/L (< 0.000125 |ig/L to 0.003 |ig/L) (D'eon et al., 2009 as cited in
NCBI, 2022). Also, PFNA concentrations ranged from 0.80 ng/L to 2.4 ng/L (0.0008 |ig/L to
0.0024 |ig/L) in surface water samples collected from Highland Creek watershed, Canada in
2010 (Meyer et al., 2011 as cited in NCBI, 2022). Concentrations of PFNA in lake water samples
collected from four lakes on Cornwallis Island, Canada from 2003 to 2005 ranged from not
detected to 6.1 ng/L (0.0061 |ig/L) (Stock et al., 2007 as cited in NCBI, 2022).
RSC for PFNA
Literature Search and Screening
In 2020, EPA conducted a broad literature search to evaluate evidence for pathways of human
exposure to eight PFAS chemicals (PFOA, PFOS, PFBA, PFBS, PFDA, perfluorohexanoic acid
(PFHxA), PFHxS, and PFNA) (Holder et al., in prep). This search was not date limited and
spanned the information collected across the Web of Science (WOS), PubMed, and
ToxNet/ToxLine (now ProQuest) databases. The results of the PFNA literature search of
publicly available sources are available through EPA's Health & Environmental Resource
Online website at https://hero.epa.eov/hero/index.cfm/proiect/paee/proiect id/2633.
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The 2,408 literature search results for PFNA were imported into SWIFT-Review (Sciome, LLC,
Research Triangle Park, NC) and filtered through the Evidence Stream tags to identify human
studies and non-human (i.e., those not identified as human) studies (Holder et al., in prep).
Studies identified as human studies were further categorized into seven major PFAS pathways
(Cleaning Products, Clothing, Environmental Media, Food Packaging, Home
Products/Articles/Materials, Personal Care Products, and Specialty Products) as well as an
additional category for Human Exposure Measures. Non-human studies were grouped into the
same seven major PFAS pathway categories, except that the Environmental Media category did
not include soil, wastewater, or landfill. Only studies published between 2003 and 2020 were
considered. Application of the SWIFT-Review tags identified 1,359 peer-reviewed papers
matching these criteria for PFNA.
Holder et al. (in prep) screened the 1,359 papers to identify studies reporting measured
occurrence of PFNA in human matrices and media commonly related to human exposure (human
blood/serum/urine, drinking water, food, food contact materials, consumer products, indoor dust,
indoor and ambient air, and soil). For this synthesis, additional screening was conducted to
identify studies relevant to surface water (freshwater only) and groundwater using a keyword13
search for water terms.
Following the Population, Exposure, Comparator, and Outcome (PECO) criteria outlined in
Table A-l, the title and abstract of each study were independently screened for relevance by two
screeners using litstream™. A study was included as relevant if it was unclear from the title and
abstract whether it met the inclusion criteria. When two screeners did not agree if a study should
be included or excluded, a third reviewer was consulted to make a final decision. The title and
abstract screening of Holder et al. (in prep) and of this synthesis resulted in 679 unique studies
being tagged as relevant (i.e., having data on occurrence of PFNA in exposure media of interest)
that were further screened with full-text review using the same inclusion criteria. After additional
review of the evidence collected by Holder et al. (in prep), 98 studies originally identified for
other PFAS also contained information relevant to PFNA. Based on full-text review, 171 studies
were identified as having relevant, extractable data for PFNA from the United States, Canada, or
Europe for environmental media, not including studies with only human biomonitoring data. Of
these 171 studies, 156 were identified from Holder et al. (in prep), where primary data were
extracted into a comprehensive evidence database. Parameters of interest included: sampling
dates and locations, numbers of collection sites and participants, analytical methods, limits of
detection and detection frequencies, and occurrence statistics. Fifteen of the 171 studies were
identified in this synthesis as containing primary data on only surface water and/or groundwater.
The evidence database of Holder et al. (in prep) additionally identified 18 studies for which the
main article was not available for review. As part of this synthesis, 17 of the 18 studies could be
retrieved. An additional three references were identified through gray literature sources that were
included to supplement the search results. The combined 20 studies underwent full-text
screening using the inclusion criteria in Table A-l. Based on full-text review, five studies were
identified as relevant.
13 Keyword list: water, aquifer, direct water, freshwater, fresh water, groundwater, ground water, indirect water,
lake, meltwater, melt water, natural water, overland flow, recreation water, recreational water, river, riverine water,
riverwater, river water, springwater, spring water, stream, surface water, total water, water supply
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Table A-l. Populations, Exposures, Comparators, and Outcomes (PECO) Criteria
PECO Element Inclusion Criteria
Population Adults and/or children in the general population and populations in the
vicinity of PFAS point sources from the United States, Canada, or Europe
Exposure Primary data from peer-reviewed studies collected in any of the following
media: ambient air, consumer products, drinking water, dust, food, food
packaging, groundwater51, human blood/serum/urine, indoor air, landfill,
sediment, soil, surface water51 (freshwater), wastewater/biosolids/sludge
Comparator Not applicable
Outcome Measured concentrations of PFNA (or measured emissions from food
packaging and consumer products only)
Note:
a Surface water and groundwater were not included as relevant media in Holder et al. (in prep). Studies were re-screened for these
two media in this synthesis.
Using the screening results from the evidence database and this synthesis, a total of 176 studies
were identified as relevant. Forty-seven of these contained information relevant to the U.S. and
were summarized for this effort.
Additional Screening
EPA also searched the following gray literature sources for information related to relative
exposure of PFNA for all potentially relevant routes of exposure (oral, inhalation, dermal) and
exposure pathways relevant to humans:
• ATSDR's Toxicological Profiles;
• CDC's national reports on human exposures to environmental chemicals;
• EPA's CompTox Chemicals Dashboard;
• EPA's fish tissue studies;
• EPA's Toxics Release Inventory;
• EPA's UCMR data;
• Relevant documents submitted under the Toxic Substances Control Act and relevant
reports from EPA's Office of Chemical Safety and Pollution Prevention;
• U.S. Food and Drug Administration's (FDA's) Total Diet Studies and other similar
publications from FDA, U.S. Department of Agriculture, and Health Canada;
• National Oceanic and Atmospheric Administration's (NOAA's) National Centers for
Coastal Ocean Science data collections;
• National Science Foundation direct and indirect food and/or certified drinking water
additives;
• PubChem compound summaries;
• Relevant sources identified in the relative source contribution discussions (Section 5) of
EPA's Proposed Approaches to the Derivation of a Draft Maximum Contaminant Level
Goal for Perfluorooctanoic Acid (PFOA)/Perfluorooctane Sulfonic Acid (PFOS) in
Drinking Water; and
• Additional sources, as needed.
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EPA has included available information from these gray literature sources for PFNA relevant to
its uses, chemical and physical properties, and for occurrence in drinking water (directly or
indirectly in beverages like coffee, tea, commercial beverages, or soup), ambient air, foods
(including fish and shellfish), incidental soil/dust ingestion, and consumer products. EPA has
also included available information specific to PFNA below on any regulations that may restrict
PFNA levels in media (e.g., water quality standards, air quality standards, food tolerance levels).
Summary of Potential Sources of PFNA Exposure
EPA presents information below from studies performed in the United States. While studies from
non-U. S. countries inform an understanding global exposure sources and trends, the RSC
determination is based on the available data for the United States.
Dietary Sources
Seafood
PFNA was detected in 108 of 157 fish tissue composite samples collected during EPA's National
Lake Fish Tissue Study, with a maximum concentration of 9.70 ng/g and a 50th percentile
concentration of 0.32 ng/g (Stahl et al., 2014). It was detected in one of 162 fish tissue composite
samples collected during EPA's 2008-2009 National Rivers and Streams Assessment (NRSA) at
a concentration of 2.48 ng/g (Stahl et al., 2014). More recently, PFNA was detected in 135 of
349 fish tissue composite samples at concentrations ranging from 0.100 ng/g to 1.910 ng/g in
EPA's 2013-2014 NRSA (EPA, 2020). PFNA was also detected in 119 of 152 fish tissue
composite samples at concentrations ranging from 0.12 ng/g to 9.32 ng/g in EPA's 2015 Great
Lakes Human Health Fish Fillet Tissue Study (EPA, 2021). In 2001, PFNA was detected at mean
concentrations of 1.0 ng/g, 0.57 ng/g, 2.8 ng/g, 2.9 ng/g, and 1.1 ng/g (wet weight) in whole
body homogenates of lake trout collected from Lake Superior, Lake Michigan, Lake Huron,
Lake Erie and Lake Ontario, respectively (Furdui et al., 2007 as cited in ATSDR, 2021 and
NCBI, 2022). In addition, PFNA was detected in lake trout at concentrations of 0.70 ng/g for
Lake Superior, 1.4 ng/g for Lake Huron, 2.6 ng/g for eastern Lake Erie, and 0.90 ng/g for Lake
Ontario; PFNA was also detected at a concentration of 1.2 ng/g in walleye collected from
western Lake Erie (De Silva et al., 2011; ATSDR, 2021). PFNA was detected in mixtures of
whole fish from the Missouri River, the Mississippi River, and the Ohio River at concentrations
of 0.43 ng/g, 0.78 ng/g, and 1.03 ng/g, respectively (Ye et al., 2008; ATSDR, 2021).
Concentrations of PFNA ranged from 0.01 ng/g to 0.73 ng/g in capelin whole body
samples, < 0.09 ng/g to 1.3 ng/g in cod muscle samples, and 0.05 ng/g to 8.0 ng/g in salmon
muscle samples collected from the Hudson Bay region of northeast Canada in 1999 to 2003
(Kelly, et al. 2009; NCBI, 2022). PFNA was not included in NOAA's National Centers for
Coastal Ocean Science, National Status and Trends Data (NOAA, 2022).
Five additional studies were identified that evaluated PFNA levels in seafood (Byrne et al., 2017;
Chiesa et al., 2019; Schecter et al., 2010; Young et al., 2013, 2022) (Table A-2). Four of these
studies analyzed fish purchased from stores and fish markets. PFNA was detected infrequently in
samples reported in Chiesa et al (2019), Schecter et al. (2010) and Young et al. (2013): one of
ten samples of striped bass (1.4 ng/g) and in one of nine samples of shrimp (1.2 ng/g), but not in
samples of crab meat, catfish, clams, cod, flounder, pangasius, pollock, tuna (including canned),
salmon, scallops, tilapia, canned sardines, or frozen fish sticks. No other fish types were sampled
in these three studies, and other than canned tuna and sardines, none were analyzed as prepared
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for eating. Seafood samples reported in Young et al. (2022) reported detectable PFNA in five out
of the eight types of seafood evaluated. These included canned clams, canned tuna, cod, crab
meat, and pollock (fish fillets and frozen fish sticks). No PFNA was detected in salmon, tilapia
or shrimp. Seafood packaging was also evaluated for PFAS coatings, and it was determined the
packaging did not contribute to any PFAS concentrations observed in the study.
One study evaluated fish samples collected directly from rivers and lakes (Byrne et al., 2017). As
part of a study to assess exposure to PFNA and other PFAS among residents of two remote
Alaska Native villages on St. Lawrence Island, Byrne et al. (2017) measured PFAS
concentrations in stickleback and Alaska blackfish, resident fish used as sentinel species to detect
accumulation of PFAS in the local environment. Stickleback were collected from three
locations—Suqitughneq (Suqi) River watershed (n = 9 composite samples), Tapisaggak (Tapi)
River (n = 2 composite samples), and Troutman Lake (n = 3 composite samples). Blackfish were
collected from the Suqi River (n = 29) but were not found in the other water bodies. Authors
reported that the Suqi River watershed was upstream and downstream of a formerly used defense
site and Tapi River was east of a military site, however at the start of the study none of the sites
were known to be contaminated with PFAS. The sample dates were not reported. PFNA was not
detected in the blackfish samples but was detected in 100%, 56%, and 50% of stickleback
samples from Troutman Lake, Suqi River, and Tapi River, respectively, with authors noting that
PFNA was the most frequently detected PFAS in stickleback. PFNA concentrations ranged
between 2.72 ng/g-4.13 ng/g ww at Troutman Lake, from below the limit of detection (LOD) to
1.52 ng/g ww at Suqi River, and < LOD-O.78 ng/g ww at Tapi River (LOD not reported; limit of
quantitation (LOQ) = 0.5-1 ng/g ww). The authors reported that total PFAS levels were
"exceptionally high" in Troutman Lake and hypothesized that stickleback were exposed to a
local PFAS source and that contaminant may be leaching from village and military landfills.
The remaining four studies purchased seafood from stores and fish markets (Chiesa et al., 2019;
Schecter et al., 2010; Young et al., 2013, 2022). Young et al. (2013) assessed fish and shellfish
collected in 2010-2012 from retail markets across the continental United States. Retail markets
in California, Florida, Illinois, Mississippi, New Jersey, New York, Tennessee, Texas, and
Washington, D.C. were represented. Authors selected the ten most consumed fish and shellfish
in the United States that were farm raised, wild caught, or had unknown origin. Among the crab
meat, shrimp, striped bass, catfish, clams, cod, flounder, pangasius, pollock, tuna, salmon,
scallops, and tilapia, PFNA was only detected in one of nine samples of shrimp at a
concentration of 1.2 ng/g and one of ten samples of striped bass at a concentration of 1.4 ng/g.
Young et al. (2022) evaluated fish and shellfish collected from retail markets in the Washington,
D.C. metropolitan area, from March 2021 through May 2022. Some clam samples were also
purchased online. Eight seafood products were selected that represented those in the top ten
types of seafood consumed in the United States. Seafood products were farm raised, wild-caught
or of unknown origin. PFNA was detected in all clam (n=10) and crab (n=l 1) samples with
concentrations ranging up to 796 ng/kg for clams and 350 ng/kg for crabs. Samples of cod (40%,
n=10), pollock (20%), n=10) and canned/pouch tuna (30%, n=10) also had detectable PFNA with
concentrations of 45-103 ng/kg, 100-106 ng/kg and 44-77 ng/kg, respectively. Salmon, shrimp
and tilapia did not have detectable levels of PFNA (MDL=30-39 ng/kg). Schecter et al. (2010)
evaluated PFNA and other PFAS in seafood collected from five Dallas, Texas grocery stores in
2009. The origin or source of seafood was not described. Seafood included canned sardines in
water, canned tuna, fresh catfish fillet, cod, frozen fish sticks, salmon, and tilapia (n = 1
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composite sample for each seafood type). PFNA was not detected in any of the seafood samples.
Finally, in a multicontinental study, Chiesa et al. (2019) collected salmon from a wholesale fish
market in Milan, Italy; the sampling year was not reported. Wild-caught salmon samples
originated from the United States (n = 7), Canada (n = 15), and Scotland (n = 2), while farmed
salmon samples originated from Norway (n = 25) and Scotland (n = 17). Among the salmon that
originated from the United States - Pacific Ocean (Food and Agriculture Organization Area
(FAO) 67 and 77), two species—Oncorhynchus kisutch and Oncorhynchus keta—were analyzed,
with PFNA not detected in either species (LOQ = 0.005 ng/g). PFNA was also not detected in
wild-caught salmon from Canada and Scotland.
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Table A-2. Summary of PFNA Data in Seafood
Study
Location and Source
Seafood Type
Results
United States
Byrne et al. (2017)
United States (Alaska)
Stickleback collected from three locations
on St. Lawrence Island: Suqitughneq (Suqi)
River watershed (upstream and downstream
of a formerly used defense site),
Tapisaggak (Tapi) River (located
approximately 5 km east of military site),
and Troutman Lake, a coastal lake situated
adjacent to the village of Gambell.
Alaska blackfish collected from the Suqi
River but were absent from the other water
bodies.
Sampling year not reported. No sites were
known to be contaminated with PFAS at
the initiation of the study.
Stickleback and Alaska
blackfish
Strickleback:
Troutman Lake: n = 3*, DFa 100%,
mean3 (range) = 3.43 (2.72-
4.13) ng/g ww
Suqi River: n = 9*, DFa 56%,
range = < LOD-1.52] ng/g ww
Tapi River: n = 2*, DFa 50%, range
= < LOD-O.78 ng/g ww
Blackfish: n = 29, DF 0%
(LOQ = 0.5-1 ng/g ww for all PFAS)
*Number of composite samples, each
composed of -10 stickleback fish
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Study
Location and Source
Seafood Type
Results
United States
Young et al. (2013)
United States (California; Illinois;
Mississippi; Tennessee; Florida; New
Jersey; New York; Texas; Washington,
DC.)
Fish and shellfish collected from retail
markets in 11 areas across the continental
United States from 2010-2012. The fish
Crab, shrimp, striped
bass, catfish, clams, cod,
flounder, pangasius,
pollock, tuna (can and
pouch), salmon, scallops
(bay and sea), tilapia
Shrimp: n = 9, DFa 11%, range = ND-
1.2* ng/g
Striped bass: n = 10, DFa 10%,
range = ND-1.4* ng/g
Crab meat: n = 1, DF 0%
Catfish: n = 13, DF 0%
and shellfish included farm raised, wild
caught, and unknown origin, as well as
freshwater fish, saltwater fish, and
euryhaline fish.
Crab meat, clams, cod, flounder, pangasius,
salmon, scallops, and tilapia purchased
from Washington, D.C. Shrimp purchased
from Orlando, Florida; Memphis,
Tennessee; and Nashville, Tennessee.
Striped bass purchased from New York,
New York and Cherry Hill, New Jersey.
Catfish purchased from Indianola,
Mississippi; Dallas, Texas; Tampa, Florida;
and Orlando, Florida. Pollock purchased
from Huntington Beach, California. Tuna
purchased from Chicago, Illinois.
Clams: n = 1, DF 0%
Cod: n = 1, DF 0%
Flounder: n = 1, DF 0%
Pangasius: n = 1, DF 0%
Pollock: n = 1, DF 0%
Tuna: n = 3, DF 0%
Salmon: n = 2, DF 0%
Scallops: n = 2, DF 0%
Tilapia: n = 1, DF 0%
(MDL = 0.60 ng/g for all seafood)
*This value was above the MDL but below
the LOQ; LOQ is estimated as 3x the MDL
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Study
Location and Source
Seafood Type
Results
United States
Young et al. (2022)
United States (Washington, D.C.)
Fish and shellfish collected from retail
markets in the Washington, D.C.
metropolitan area from 2021-2022. Fish
and shellfish samples included farm raised,
wild-caught and unknown origin. Country
of origin was provided, if known.
Ten samples of each seafood type, except
for crab, which included 11 samples.
Crab, clams (can),
shrimp, cod, pollock (fish
sticks, fillet), salmon,
tuna (can and pouch),
tilapia
Clam meat: n=10, DF 100% range=333-
796 ng/kg
Crab: n=ll, DF 100%, range=54-350
ng/kg
Cod: n=10, DF 40%, range=
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Study
Location and Source
Seafood Type
Results
United States
Chiesa et al. (2019) United States (Pacific Ocean) Wild-caught salmon Oncorhynchus kisutch: n = 5, DF 0%
Wild-caught fish were collected at a (Oncorhynchus kisutch Oncorhynchus keta: n = 2, DF 0%
wholesale fish market in Milan, Italy. an<^ Oncorhynchus keta) (LOQ = 0 005 ng/g)
Sampling year was not reported. The wild-
caught salmon were from USA-Pacific
Ocean (Food and Agriculture Organization
Area 67 and 77).
Notes: DF = detection frequency; ww = wet weight, LOD = limit of detection; LOQ = limit of quantitation; MDL = method detection limit; ND = not detected.
a The DF and/or mean was not reported in the study and was calculated in this synthesis. Means were calculated only when DF = 100%.
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Other Food Types
PFNA was included in a suite of PFAS evaluated in FDA's 2019, 2021, and 2022 Total Diet
Study Sampling (U.S. FDA, 2020a,b, 2021a,b, 2023a,b,c,d); it was detected at concentrations of
233 ng/kg (0.233 ng/g) in baked cod and 50 ng/kg (0.050 ng/g) in frozen (oven-cooked) fish
sticks or patties in 2021, but it was not detected in any of the other food samples tested. It should
be noted that FDA indicated that the sample sizes used in the PFAS 2019, 2021, and 2022 Total
Diet Study Sampling were limited and that the results should not be used to draw definitive
conclusions about PFAS levels in the general food supply (U.S. FDA, 2022c). PFNA was not
detected in milk samples collected from a farm with groundwater known to be contaminated with
PFAS; however, it was detected in produce (corn) collected from an area near a PFAS
production plant in FDA studies of the potential PFAS exposure to the U.S. population (U.S.
FDA 2018, 2021c). PFNA was detected in beef steak in the Canadian Total Diet studies from
1992 to 2004, but it was not detected in any of the other food samples tested (ATSDR, 2021;
Tittlemier et al., 2007).
Seven U.S. studies were identified that examined PFNA in breastmilk or food types other than
breastmilk (Blaine et al., 2013, 2014; Kuklenyik et al., 2004; Schecter et al., 2010; Tipton et al.,
2017; von Ehrenstein et al., 2009; Young et al., 2012) (Table A-3). Few U.S. studies analyzed
foods from any one origin—only two studies sampled from store- or market-bought meats, eggs,
produce, and dairy, one studied wild alligator meat, two sampled from crops (produce and corn
grain and stover) grown in biosolids-amended soils (and also control and municipal soils) as part
of greenhouse and field studies, and two studied breastmilk.
Two studies purchased food items from stores and markets for evaluation (Schecter et al., 2010;
Young et al., 2012). Schecter et al. (2010) assessed PFNA and other PFAS in food samples
collected from five Dallas, Texas grocery stores in 2009. The origin or source of each food was
not described. Food items included meat products (bacon, canned chili, chicken breast, ground
beef, roast beef, ham, sausage, and turkey), dairy (butter, cheeses, frozen yogurt, ice cream, milk,
and yogurt), eggs, and grains (cereal), fruits and vegetables (apples, potatoes), and fats/other
(canola oil, margarine, olive oil, peanut butter). PFNA was not detected in any of the food
samples. In Young et al. (2012), cow milk was purchased from retail markets across the
continental United States representing 17 states; the sampling year was not reported. Cow milk
samples included organic milk, vitamin D added milk, and ultra-pasteurized milk. PFNA was not
detected in any of the 49 retail milk samples (method detection limit (MDL) = 0.28 ng/g).
One study investigated PFAS levels from wild meat (Tipton et al., 2017). Tipton et al. (2017)
assessed alligator tail meat that was collected during the South Carolina recreational hunting
season between September to October 2015. Tail meat samples were collected from four
different public hunt units—Southern Coastal, Middle Coastal, Midlands, and Pee Dee. PFNA
was detected in samples from all hunt units with the exception of the Midlands (n = 2), where
PFNA was not detected. Median concentrations from Southern Coastal (n = 19), Middle Coastal
(n = 17), and Pee Dee (n = 2) were 0.107 ng/g, 0.102 ng/g, and 0.117 ng/g wet mass,
respectively.
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Two studies by Blaine et al. (2013, 2014) evaluated PFNA in crops grown in greenhouse and
field studies. In Blaine et al. (2014), PFAS levels were measured in celery root, pea fruit, and
radish root grown in a greenhouse study with control (unamended) soil, industrially impacted
soil, and municipal soil (n = 3-5). PFNA was detected in radish root and celery shoot from all
three soils and pea fruit from only industrially impacted soil. Mean concentrations of PFNA in
radish root for the control, industrially impacted, and municipal soil were 4.79 ng/g, 26.88 ng/g,
and 5.99 ng/g, respectively. Mean concentrations of PFNA in celery shoot for the control,
industrially impacted, and municipal soil were 1.89 ng/g, 13.81 ng/g, and 1.62 ng/g, respectively.
The mean concentration of PFNA in pea fruit in the industrially impacted soil was 1.45 ng/g.
Authors noted minor cross-contamination of the control soil due to the proximity of the
unamended soil to biosolids-amended soil. In Blaine et al. (2013), authors studied the uptake of
PFAS into edible crops in both field and greenhouse studies. In the field study, PFAS levels were
measured in corn grain and corn stover grown with control (unamended), urban biosolids-
amended, and rural biosolids-amended soil (n = 3-7). Mean PFNA concentrations were below
the LOQ in both corn grain and corn stover grown in any field study plots (<0.10 ng/g for corn
grain; < 0.29 ng/g for corn stover). In the greenhouse study, lettuce and tomato plants were
grown in control soil, industrially impacted soil, or municipal soil (n = 3-5). Mean PFNA
concentrations were below the LOQ (2.96 ng/g) in any tomato plants but was detected in lettuce
grown in industrially impacted soil and municipal soil at mean concentrations of 57.39 ng/g and
4.73 ng/g, respectively. PFNA was not detected above the LOQ (0.04 ng/g) in lettuce grown in
control soil. Sampling year was not reported.
The remaining two studies evaluated the occurrence of PFNA in breastmilk (Kuklenyik et al.,
2004; von Ehrenstein et al., 2009). von Ehrenstein et al. (2009) collected breastmilk samples
between December 2004 and July 2005 from women between the ages of 18 and 38 at the time
of recruitment as part of the pilot study Methods Advancement for Milk Analysis (MAMA).
Women provided milk samples at two visits—the first visit was 2-7 weeks postpartum, and the
second visit was 3-4 months postpartum. PFNA was not detected in any of the samples from the
first visit (n = 18) or second visit (n = 20). Similarly, PFNA was below the LOD (1.0 ng/mL) in
the samples reported by Kuklenyik et al. (2004). Kuklenyik et al. (2004) did not report
information on the breastmilk donors or the sampling procedure as it was unavailable; PFNA
was not detected in either of the two samples.
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Table A-3. Summary of PFNA Data in Other Food
Study Location and Source
Food Types
PFNA Results
United States
Schecter et al. (2010) United States (Texas)
Food samples from five different
grocery stores in Dallas, Texas were
collected in 2009. Ten individual
samples were collected for each food
type and combined to form composite
samples. The origin/source of the food
Dairy; fruits and
vegetables; grains;
meat; fats/other
Meat
Hamburger: n = 1, DF 0%
Bacon: n = 1, DF 0%
Sliced turkey: n = 1, DF 0%
Sausages: n = 1, DF 0%
Ham: n = 1, DF 0%
Sliced chicken breast: n = 1, DF 0%
samples were not reported. Roast beef: n 1, DF 0%
Canned chili: n = 1, DF 0%
Dairy and Eggs
Butter: n = 1, DF 0%
American cheese: n = 1, DF 0%
Other cheese: n = 1, DF 0%
Whole milk: n = 1, DF 0%
Ice cream: n = 1, DF 0%
Frozen yogurt: n = 1, DF 0%
Whole milk yogurt: n = 1, DF 0%
Cream cheese: n = 1, DF 0%
Eggs: n = 1, DF 0%
Grains
Cereals: n = 1, DF 0%
Fruits and Vegetables
Apples: n = 1, DF 0%
Potatoes: n = 1, DF 0%
Fats and Other
Olive oil: n = 1, DF 0%
Canola oil: n = 1, DF 0%
Margarine: n = 1, DF 0%
Peanut butter: n = 1, DF 0%
(LOD not reported for any food item)
*n reflects number of composite samples, each
composed of-10 individual samples
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Study
Location and Source
Food Types
PFNA Results
United States
Young et al. (2012)
United States (17 states)
Retail cow's milk samples were all
pasteurized whole milk, commercially
available, and purchased at retail
markets across the continental United
States representing 17 states. Samples
were organic milk, vitamin D added
milk, or ultra-pasteurized milk.
Sampling year not reported.
Dairy
n = 49, DF 0%
(MDL = 0.28 ng/g)
Tipton etal. (2017)
United States (South Carolina)
Alligator tail meat samples were
collected from a local wild game meat
processer during the South Carolina
recreational hunt season between
September to October 2015. Samples
were from four different public hunt
units—Southern Coastal, Middle Coast,
Midlands, and Pee Dee.
Meat
Alligator tail:
Southern coastal: n = 19, DFa 74%, median
(range) = 0.107 (< 0.088-0.551) ng/g wet mass
Middle coastal: n = 17, DFa 65%, median
(range) = 0.102 (< 0.073-0.553) ng/g wet mass
Pee Dee: n = 2, DFa 100%, median
(range) = 0.117 (0.100-0.135) ng/g wet mass
Midlands: n = 5, DF 0%
(RL not reported)
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Study
Location and Source
Food Types
PFNA Results
United States
Blaine et al. (2014)
United States (Midwest)
Crops grown in in greenhouse study
with control (unamended), industrially
impacted soil, or municipal soil.
Control soil had minor cross-
contamination due to proximity to
biosolids-amended fields. Industrially
impacted soil was amended with
industrially impacted biosolids, and
municipal soil was amended with
municipal biosolids for over 20 years.
Crops grown in the greenhouse study
were grown from seed in pots, which
were randomly arranged within the
greenhouse. Sampling year not
reported.
Fruits and vegetables Radish root:
Control: n = 3-5, DF NR, mean = 4.79 ng/g
Industrially impacted; n = 3-5, DF NR,
mean = 26.68 ng/g
Municipal: n = 3-5, DF NR, mean = 5.99 ng/g
Celery shoot:
Control: n = 3-5, DF NR, mean = 1.89 ng/g
Industrially impacted: n = 3-5, DF NR,
mean = 13.81 ng/g
Municipal: n = 3-5, DF NR, mean = 1.62 ng/g
Pea fruit:
Control: n = 3-5, DF 0%
Industrially impacted: n = 3-5, DF NR,
mean = 1.45 ng/g
Municipal: n = 3-5, DF 0%
(LOQ = 0.07 ng/g)
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Study
Location and Source
Food Types
PFNA Results
United States
Blaine et al. (2013)
United States (Midwest)
Crops grown in urban and rural full-
scale field study with control
(unamended) and biosolids-amended
soil. Three agricultural fields were
amended (0.5x, lx, or 2*) with
municipal biosolids. Urban biosolids
(1 x and 2x) were from a WWTP
receiving both domestic and industrial
waste. Rural biosolids (0.5x) were from
a WWTP receiving domestic waste
only. Control plots were proximal to the
rural and urban amended corn grain and
corn stover field sites; sampling year
not provided.
Crops grown in greenhouse study with
control (nonamended) and biosolids-
amended soil. Nonamended soil
obtained from a field that received
commercial fertilizers and had a similar
cropping system as the nearby
municipal soil site. Municipal soil was
obtained from a reclamation site in
Illinois where municipal biosolids were
applied at reclamation rates for 20
years, reaching the cumulative biosolids
application rate of 1,654 Mg/ha.
Industrially impacted soil was created
by mixing composted biosolids from a
small municipal (but impacted by
PFAA manufacturing) WWTP with
control soil on a 10% mass basis.
Sampling year not provided.
Fruits and vegetables;
grains
Field study:
Corn grain:
Urban nonamended: n = 3-7, DF NR,
mean = < 0.10 ng/g
Urban 1 x: n = 3-7, DF NR, mean = < 0.10 ng/g
Urban 2x: n = 3-7, DF NR, mean = < 0.10 ng/g
Rural nonamended: n = 3-7, DF NR,
mean = < 0.10 ng/g
Rural 0.5 x: n = 3-7, DF NR, mean = <0.10 ng/g
Corn stover:
Urban nonamended: n = 3-7, DF NR,
mean = < 0.29 ng/g
Urban 1 x; n = 3-7, DF NR, mean = < 0.29 ng/g
Urban 2x: n = 3-7, DF NR, mean = < 0.29 ng/g
Rural nonamended: n = 3-7, DF NR,
mean = < 0.29 ng/g
Rural 0.5 x: n = 3-7, DF NR, mean = < 0.29 ng/g
(LOQ = 0.10 ng/g for corn grain; LOQ = 0.29 ng/g
for corn stover)
Greenhouse study:
Lettuce:
Nonamended: n = 3-5, DF NR, mean = < 0.04 ng/g
Industrially impacted: n = 3-5, DF NR,
mean = 57.39 ng/g
Municipal: n = 3-5, DF NR, mean = 4.73 ng/g
Tomato:
Nonamended: n = 3-5, DF NR, mean = < 2.86 ng/g
Industrially impacted: n = 3-5, DF NR,
mean = < 2.86 ng/g
Municipal: n = 3-5, DF NR, mean = < 2.86 ng/g
(LOQ = 0.04 ng/g for lettuce; LOQ = 2.86 ng/g for
tomato)
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Study
Location and Source
Food Types
PFNA Results
United States
von Ehrenstein et al.
United States (North Carolina)
Breastmilk
Visit #1: = 18, DF 0%
(2009)
As part of the Methods Advancement
for Milk Analysis (MAMA) pilot study,
34 breastfeeding women aged 18 to
38 years at recruitment provided
breastmilk samples at two visits. The
first visit occurred 2-7 weeks
Visit #2: n = 20, DF 0%
(LOQ = 0.30 ng/mL)
postpartum, and the second visit
occurred 3-4 months postpartum. Both
visits were between December 2004
and July 2005.
Kuklenyik et al. United States (Georgia) Breastmilk n = 2, DF 0%
(2004) Authors reported that no information (LOD =1.0 ng/mL)
was provided on the human milk donors
or the sampling procedure.
Notes: DF = detection frequency; LOD = limit of detection; LOQ = limit of quantitation; 0.5 x, 1*, or 2* = Vi, 1, or 2 times the agronomic rate of biosolids application to meet
nitrogen requirements of the crop; MDL = method detection limit; NR = not reported; RL = reporting limit; WWTP = wastewater treatment plant.
Bold indicates detected levels of PFNA in food.
a The DF and/or mean was not reported in the study and was calculated in this synthesis. Means were calculated only when DF = 100%.
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Food Contact Materials
PFNA has been detected in the packaging paper of one of three brands of microwave popcorn
bags (uncooked and cooked) (Sinclair et al., 2007; ATSDR, 2021). One other study analyzed the
occurrence of PFAS in treated food contact paper and other consumer products purchased from
local retailers and online stores in the United States between March 2007 and September 2011
(Liu et al., 2014). All treated food contact paper was manufactured in the United States. PFNA
was detected in 33% of samples (n = 9), with two of the detects below 10 ng/g and the third
detect at 212 ng/g.
In 2011, FDA reached a voluntary agreement with industry to remove from the market certain
PFAS greaseproofing agents used in fast food packaging. As such, the reported detection of
PFNA in fast food packaging in the above cited studies may be an overestimation of the
occurrence and levels of PFNA in current food packaging paper.
Consumer Products
Since the 1950s, PFNA has been used in industrial and consumer products, including fabric and
carpet protective coatings, paper coatings, insecticide formulations, and surfactants (NCBI,
2022). PFNA and other long-chain PFAS are found in aqueous film forming foams, cosmetics,
dental floss, floor polish, leather, food packaging materials, lithium batteries, ski wax, treated
apparel, work apparel for medical staff, pilots, and firefighters, and in hair treatment products
(NCBI, 2022).
Based on limited testing, PFNA has been detected in rinsates from fluorinated high-density
polyethylene (HDPE) containers used by one pesticide product supplier (EPA, 2022a). PFNA is
not a registered pesticide under the Federal Insecticide, Fungicide, and Rodenticide Act
(FIFRA), and EPA does not set a 40 CFR Part 180 pesticide tolerance in food and feed
commodities for PFNA (U.S. GPO, 2022). Maximum residue levels for PFNA were not found in
the Global Maximum Residue Level Database (Bryant Christie Inc., 2022).
Two studies were identified that analyzed PFNA concentrations in a range of consumer products,
including children's nap mats, household carpet/fabric-care liquids, and textiles (Liu et al., 2014;
Zheng et al., 2020) (Table A-4). Of the two U.S. studies, the consumer products evaluated are
likely used by adults (e.g., floor waxes), can come into contact with both adults and children
(e.g., treated upholstery), or the user was not specified (e.g., clothing).
Zheng et al. (2020) determined the occurrence of ionic and neutral PFAS in the childcare
environment (dust and nap mats). Samples of children's nap mats were collected from seven
Seattle childcare centers (n = 26; 20 polyurethane foam, 6 vinyl cover samples). PFNA was
detected in 36% of nap mat samples with a mean concentration of 0.19 ng/g. Half of the
analyzed mats were purchased as new products and the other half were used. The authors
reported that total PFAS levels in the new vs. used mats were not significantly different. Total
PFAS levels in mat foam vs. mat covers were also similar. Based on these results, the authors
suggested that indoor air was not the major source of PFAS in mats and that PFAS in mats could
be associated with the manufacturing process. Liu et al. (2014) analyzed the occurrence of PFAS
in consumer products (including pre-treated carpeting, commercial carpet-care liquids,
household carpet/fabric-care liquids, treated apparel, treated home textiles and upholstery,
treated non-woven medical garments, treated floor waxes and stone-wood sealants, membranes
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for apparel, and thread-sealant tapes and pastes) purchased between March 2007 and September
2001 from local retailers and online stores in the United States. The consumer products
originated from the United States, England, Vietnam, China, Thailand, El Salvador, Bangladesh,
Dominican Republic, Malaysia, and Indonesia. PFNA was detected in 44% of 9 pre-treated
carpeting samples (ranging from below the detection limit (BDL) to 236 ng/g); in 58% of 12
commercial carpet/fabric-care liquid samples (BDL-8,860 ng/g); in 15% of 13 household
carpet/fabric-care liquid and foam samples (BDL-37.3 ng/g); in 60% of 15 treated apparel
samples (BDL-235 ng/g); in 100% of 6 treated home textile and upholstery samples with a mean
of 42.6 ng/g; in 56% of 9 treated non-woven medical garment samples (BDL-334 ng/g); in 88%
of 8 treated floor wax and stone/wood sealant samples (BDL-2,740 ng/g); and in 75% of 8
membranes for apparel samples (BDL-12.8 ng/g). PFNA was not detected in thread-sealant
tapes and pastes (n = 6). Detection limits were not reported in the study.
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Table A-4. Summary of PFNA Consumer Product Data
Study
Location
Site Details
Results
United States
Zheng et al. (2020)
United States (Seattle,
Children's nap mat samples (n = 26, finely
N = 26, DF 36%, mean, median (range) = 0.19,
Washington)
cut) from seven Seattle childcare centers,
0.11 (ND-0.65) ng/g
including polyurethane foam (n = 20) and
(MDL = 0.08 ng/g)
vinyl cover (n = 6) samples. Sampling year
not reported.
Liu et al. (2014)
United States
(unspecified)
Consumer products commonly used
indoors were purchased between March
2007 and September 2011 from local
retailers and online stores in the United
States. The samples analyzed for PFCAs
included pre-treated carpeting, commercial
carpet/fabric-care liquids, household
carpet/fabric-care liquids and foams,
treated apparel, treated home textile and
upholstery (i.e., mattress pads), treated
non-woven medical garments, treated floor
waxes and stone-wood sealants,
membranes for apparel, and thread-sealant
tapes and pastes. The products originated
from the United States, England, Vietnam,
China, Thailand, El Salvador, Bangladesh,
Dominican Republic, Malaysia, and
Indonesia.
Pre-treated carpeting: n = 9, DFa 44%,
range = BDL-236 ng/g
Commercial carpet/fabric-care liquids: n = 12,
DFa 58%, range = BDL-8,860 ng/g
Household carpet/fabric-care liquids and foams:
n = 13, DFa 15%, range = BDL-37.3 ng/g
Treated apparel: n = 15, DFa 60%,
range = BDL-235 ng/g
Treated home textile and upholstery: n = 6, DFa
100%, mean3 (range) = 42.6 (3.80-213) ng/g
Treated non-woven medical garments: n = 9,
DFa 56%, range = BDL-334 ng/g
Treated floor waxes and stone-wood sealants:
n = 8, DF 88%, range = BDL-2,740 ng/g
Membranes for apparel: n = 8, DFa 75%,
range = BDL-12.8 ng/g
Thread-sealant tapes and pastes: n = 6, DFa 0%
(DL not reported)
Notes: BDL = below detection limit; DF = detection frequency; DL = detection limit; MDL = method detection limit; ND = not detected.
a The DF and/or mean was not reported in the study and was calculated in this synthesis. Means were calculated only when DF = 100%.
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Indoor Dust
In a Wisconsin Department of Health Services study, Knobeloch et al. (2012) examined levels of
16 perfluoroalkyl chemicals in vacuum cleaner dust from 39 Wisconsin homes across 16
counties in March and April 2008 (Table A-5). Samples from these homes built between 1890
and 2005 were collected during a pilot study to assess residential exposure to persistent
contaminants found in the Great Lakes Basin. PFNA was found in all samples at a median
concentration of 12 ng/g. The number of rooms with synthetic, wall-to-wall carpeting and the
square footage of the homes were both significantly positively correlated with dust
concentrations of PFNA. Based on the results of this study, the authors suggested that
perfluoroalkyl chemicals may be ubiquitous contaminants in U.S. homes. In an EPA study of
112 indoor dust samples collected from vacuum cleaner bags from homes and daycare centers in
North Carolina and Ohio in 2000-2001 (EPA's Children's Total Exposure to Persistent
Pesticides and Other Persistent Organic Pollutants (CTEPP) study), samples were collected from
102 homes and 10 daycare centers in North Carolina (49 homes, 5 daycare centers) and Ohio (53
homes, 5 daycare centers) (Strynar and Lindstrom, 2008). Results were not reported separately
for homes and daycares. Overall, PFNA was detected in 42.9% of all samples (n = 112) with
mean and median concentrations of 22.1 ng/g and 7.99 ng/g, respectively. The authors concluded
that the study measured perfluorinated compounds in house dust at levels that may represent an
important pathway for human exposure.
Additional peer-reviewed studies were identified that evaluated the occurrence of PFNA and
other PFAS in dust of indoor environments, primarily in homes, as well as in schools, childcare
facilities, offices, and vehicles (Byrne et al., 2017; Fraser et al., 2013; Karaskova et al., 2016;
Kato et al., 2009; Scher et al., 2019; Wu et al., 2014; Zheng et al., 2020;) (Table A-5). For those
studies with results stratified for U.S. homes, PFNA levels and detection frequencies were lowest
in a study of remote Alaska Native villages (35% detection, median below 0.2 ng/g), while in
other U.S. locations, PFNA was detected in at least 65% of samples (some studies reporting
100%) detection) at widely varying mean and median levels across the studies (from
approximately 4 ng/g to 70 ng/g). Few studies sampled childcare centers, vehicles, and offices,
and none of the reviewed studies reported measurements in other microenvironments (e.g.,
public libraries, universities).
Several studies reported results from dust samples collected only from homes (Byrne et al., 2017;
Scher et al., 2019; Wu et al., 2014), with one study sampling from locations near a PFAS
production facility. Scher et al. (2019) evaluated indoor dust in 19 homes in Minnesota within a
groundwater contamination area (GCA) in the vicinity of a former 3M PFAS production facility.
Homes within the GCA had previous or ongoing PFAS contamination in drinking water and
were served by the Oakdale, Minnesota PWS or a private well previously tested and shown to
have detectable levels of PFOA or PFOS. In the house dust samples, collected from July to
September 2010, the detection frequencies for PFNA were 68%> and 95%> for entryways to the
yard and interior living spaces such as the family or living rooms, respectively (n = 19 each),
with median concentrations of 9.7 ng/g and 26 ng/g, respectively. PFAS concentrations in both
sampling locations were higher than corresponding soil concentrations, suggesting that interior
sources were the main contributors to PFAS in house dust.
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Byrne et al. (2017) assessed exposure to PFNA and other PFAS among residents of two remote
Alaska Native villages on St. Lawrence Island. PFAS concentrations were measured in dust
collected from the surfaces of floors and furniture of 49 homes on St. Lawrence Island during
February-April of 2013 and 2014. Residents were asked not to sweep or dust for one week prior
to sampling. The authors described the overall PFAS levels in dust samples as "on the lower end
of those reported worldwide in other studies." PFNA was detected in 35% of all samples (n = 49)
with a median value below the LOD (0.1 ng/g-0.2 ng/g). Wu et al. (2014) measured
concentrations of five PFCs in residential dust in California in 2008-2009. Dust samples were
collected from the carpet or area rug in the main living area of the home. Homes of parents with
young children and homes with older adults were differentiated to characterize the relationship
between serum concentrations of PFCs and several other factors, including PFC concentrations
in residential dust. PFNA was detected in 65% of samples from households with young children
in Northern California (n = 82), with mean and median concentrations of 67.4 ng/g and
9.70 ng/g, respectively. PFNA was detected in 72% of samples from households of older adults
in central California (n = 42), with mean and median concentrations of 58.5 ng/g and 11.85 ng/g,
respectively.
Apart from the information reported by Strynar and Lindstrom (2008), one other study included
childcare centers in the locations sampled (Zheng et al., 2020). Zheng et al. (2020) collected dust
samples from seven childcare centers in Seattle, Washington (n = 14) and one childcare facility
in West Lafayette, Indiana (n = 6 across six rooms); the sampling year was not reported. The
included childcare facilities consisted of several building types, including multiple classrooms, a
former church, and a former home. Because centers were vacuumed and mopped daily, dust
samples were obtained from elevated surfaces (shelving, tops of bookcases/storage cubbies)
along with floor dust. PFNA was detected in all samples at mean and median concentrations of
3.2 ng/g and 1.7 ng/g, respectively.
One study evaluated PFNA levels in vehicles and offices, in addition to homes. Fraser et al.
(2013) collected dust samples between January and March 2009 from 3 microenvironments of 31
individuals in Boston, Massachusetts (offices (n = 31), homes (n = 30), and vehicles with
sufficient dust for analysis (n = 13)). Study participants worked in separate offices located across
seven buildings, which were categorized as Building A (n = 6), Building B (n = 17), or Other
(n = 8). Building A was a newly constructed (approximately one year prior to study initiation)
building with new carpeting and new upholstered furniture in each office; Building B was a
partially renovated (approximately one year prior to study initiation) building with new carpeting
throughout hallways and in about 10% of offices. The other buildings had no known recent
renovation occurred. Study offices were not vacuumed during the sampling week and
participants were asked not to dust or vacuum their homes and vehicles for at least one week
prior to home sampling. PFNA was detected in 94%, 67%, and 85% of office, home, and vehicle
dust samples, respectively, with geometric mean concentrations of 63.0 ng/g, 10.9 ng/g, and
14.7 ng/g, respectively. Geometric mean PFNA concentrations were statistically significantly
higher in offices compared to homes and vehicles. The study also observed that PFNA
concentration in house dust was significantly predictive of PFNA serum concentration.
Two studies evaluated dust samples collected across multiple continents (Karaskova et al., 2016;
Kato et al., 2009). Karaskova et al. (2016) examined PFAS levels in house dust collected
between April and August 2013 from the living rooms and bedrooms of 14 homes in the United
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States, 15 homes in Canada, and 12 homes in the Czech Republic (locations unspecified). PFNA
was detected in all U.S. samples (n = 20) at mean and median concentrations of 10.9 ng/g and
3.9 ng/g, respectively. The authors reported PFNA concentrations were significantly higher in
North America compared to the Czech Republic, which they indicated may suggest a faster shift
from long chain PFAS to their shorter chain homologues in Europe than in North America.
Overall, no significant differences in total PFAS concentrations were found between the
bedroom and living room in the same household although significant relationships were found
based on type of floors, number of residents, and age of the house. A second multicontinental
study (Kato et al., 2009) measured PFC concentrations in 39 household dust samples collected in
2004 from homes in the United States (Atlanta, GA) (n = 10), United Kingdom (n = 9), Germany
(n = 10), and Australia (n = 10). Across all 39 homes, PFNA was detected in 25.6% of samples
with a median concentration below the LOQ (2.6 ng/g). The authors did not report stratified
PFNA data by country.
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Table A-5. Summary of PFNA Indoor Dust Data
Study
Location
Site Details
Results
United States
Scheretal. (2019)
United States (Twin
Cities metropolitan
region, Minnesota)
Nineteen homes in three cities within a GCA near
former 3M PFAS production facility as well as from
three homes in the Twin Cities Metro outside the GCA.
Dust samples collected from an entryway to the yard
and from an interior living space (e.g., family room,
living room) in each home in July-September 2010.
Homes within the GCA had previous or ongoing PFAS
contamination in drinking water and were served by
the Oakdale, Minnesota public water system or a
private well previously tested and shown to have
detectable levels of PFOA or PFOS. Results were not
reported for homes outside the GCA.
Entryway: n = 19, DF 68%, median
(range) = 9.7 ( < RL-1,000) ng/g
Living room: n = 19, DF 95%, median
(range) = 26 ( < RL-450) ng/g
(RL = 5 ng/g)
Byrne et al. (2017)
United States (St.
Lawrence Island,
Alaska)
Dust samples collected from the surfaces of floors and
furniture from 49 homes during February-April of
2013 and 2014. Participants were asked not to sweep
or dust for one week prior to sampling.
n = 49, DF 35%, median (95th
percentile) = < LOD (1.93) ng/g
(MDL = 0.1-0.2 ng/g for all PFAS)
Wu et al. (2014)
United States (Central
Valley area, California)
Distributions of PFC dust concentrations were
determined for households with young children in
Northern California (n = 82) and households of older
adults in central California (n = 42). Dust samples were
collected in 2008-2009 from the carpet or area rug in
the main living area of the homes. Homes of parents
with young children and homes with older adults were
differentiated to characterize the relationship between
serum concentrations of PFCs and PFC concentrations
measured in residential dust.
Parents of young children: n = 82, DF
65%, mean, median (range) = 67.4,
9.70 (ND-1,910) ng/g
Older adults: n = 42, DF 72%, mean,
median (range) = 58.5, 11.85 (ND-
883) ng/g
(LOD = 0.10 ng/mL)
*Data below LOQ replaced by LOD/V2
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Study
Location
Site Details
Results
United States
Knobeloch et al.
United States (Great
Dust samples were collected by the Wisconsin
n = 39, DF 100%, median (range) = 12
(2012)
Lakes Basin, Wisconsin)
Department of Health Services from 39 Wisconsin
(1.3-280) ng/g
homes across 16 counties in March-April 2008.
(RL = 1 ng/g)
Vacuum bags were collected or bagless vacuums were
emptied into sterilized glass jars. Homes were built
between 1890 and 2005.
Zheng et al. (2020) United States (Seattle,
Washington; West
Lafayette, Indiana)
Seven childcare centers in Seattle (14 samples) and one
center in Lafayette (6 samples); sampling year not
reported. Since all centers were vacuumed and mopped
daily, dust samples from elevated surfaces (shelving,
tops of bookcases/storage cubbies) were collected
along with floor dust in the same sample.
n = 20; DF 100%, mean, median
(range) = 3.2, 1.7 (0.11-13) ng/g
(MDL = 0.08)
Strynar and United States (North
Lindstrom (2008) Carolina; Ohio)
Dust samples from vacuum cleaner bags were obtained
in 2000-2001 during the EPA's Children's Total
Exposure to Persistent Pesticides and Other Persistent
Organic Pollutants (CTEPP) study from North
Carolina (49 homes, 5 daycare centers) and Ohio (53
homes, 5 daycare centers). Vacuum cleaner bags were
only collected if available at each site.
n = 112; DF 42.9%, mean, median
(maximum) = 22.1, 7.99 (263) ng/g
(LOQ= 11.3 ng/g)
*Values below the LOQ assigned a
value of LOQ/V2
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Study
Location
Site Details
Results
United States
Fraser et al. (2013)
United States (Boston,
Massachusetts)
Dust samples were collected in January-March 2009
from offices (n = 31), homes (n = 30), and vehicles
(n = 13) of 31 individuals. Study participants worked
in separate offices located across seven buildings,
which were categorized into Building A, Building B,
and Other. Six samples were collected from Building
A, a newly constructed (approximately one year prior
to study initiation) building with new carpeting and
new upholstered furniture in each office. Seventeen
samples were collected from Building B, a partially
renovated (approximately one year prior to study
initiation) building with new carpeting throughout
hallways and in about 10% of offices. Eight samples
were collected from the other five remaining buildings
where no known recent renovation occurred. Study
offices were not vacuumed during the sampling week
and homes and vehicles were not vacuumed for at least
one week prior to sampling. Entire accessible floor
surface areas and tops of immovable furniture were
vacuumed in offices and the main living area of homes.
Entire surface areas of the front and back seats of
vehicles were vacuumed.
Number of home dust samples was reduced to 30
because 1 participant lived in a boarding house with no
main living area. Sufficient mass of dust for analysis
was available from only 13 vehicles.
Homes: n = 30, DF 67%, GM
(range) = 10.9 (6.21-1,420 ng/g)
Offices: n = 31, DF 94%, GM
(range) = 63.0 (10.9-639) ng/g
Vehicles: n = 13, DF 85%, GM
(range) = 14.7 (4.95-101 ng/g)
(LOQ = 5 ng/g)
*GM calculated by replacing
values < LOQ with LOQ/V2
* Range of detected values reported
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Study
Location
Site Details
Results
United States
Karaskova et al.
(2016)
United States
(unspecified), Canada
(unspecified), Czech
Republic (unspecified)
Fifty-six dust samples from 14 homes in the United
States, 15 homes in Canada, and 12 homes in the
Czech Republic were collected between April and
August 2013. Samples were collected in living rooms
and bedrooms.
United States: n = 20, DF 100%, mean,
median (range) = 10.9, 3.9 (1.1-
62.9) ng/g
Canada: n = 20, DF 95.0%, mean,
median (range) = 19.4, 4.4 ( < MQL-
195) ng/g
Czech Republic: n = 16, DF 50.0%,
mean, median (range) = 3.0, < MQL
(ND-11.0) ng/g
(MDL = 0.27-1.33 ng/g; MQL = 0.72-
3.48 ng/g; ranges represent lower bound
and upper bound which were calculated
by dividing the MDL/MQL by the
biggest and smallest dust sample
weight, respectively)
*Mean calculated only from
values > MQL
* Median calculated by replacing
values < MQL with V2/2*MQL
Kato et al. (2009)
United States (Atlanta,
Georgia), Germany
(unspecified), United
Kingdom (unspecified),
Australia (unspecified)
Thirty-nine household dust samples from the United
States (n = 10), Germany (n = 10), United Kingdom
(n = 9), and Australia (n = 10) collected in 2004 for
method validation. Dust sampling procedures not
described.
n = 39, DF 25.6%, median
(range) = < LOQ (< LOQ-832) ng/g
(LOQ = 2.6 ng/g)
Notes:
GCA = groundwater contamination area; DF = detection frequency; RL = reporting limit; LOD = limit of detection; MDL = method detection limit; ND = not detected;
LOQ = limit of quantitation; GM = geometric mean; MQL = method quantification limit.
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Air
Perfluoroalkyl chemicals have been released to air from wastewater treatment plants, waste
incinerators, and landfills (EPA, 2016), though there is limited information on the detection
levels or frequencies of PFNA in either indoor or ambient air. ATSDR (2021) notes
perfluoroalkyl chemicals have been detected in air and they can be transported long distances via
the atmosphere. For example, in a study performed from April 2007 to January 2009, PFNA was
detected at an average concentration of 0.3 pg/m3 in 8% of 141 atmospheric samples from
Atlantic and Southern Oceans and coastal areas of the Baltic Sea (Dreyer et al, 2009; NCBI,
2022). PFNA is not expected to be broken down directly by photolysis (NCBI, 2022). PFNA can
undergo hydroxylation in the atmosphere, with a (predicted average) atmospheric hydroxylation
rate of 8.41 x 10"13 cm3/molecule - second to a (derived) rate of 5.2 x 10"11 cm3/molecule -
second (with corresponding estimated half-life of 31 days for this reaction in air) (NCBI, 2022,
EPA, 2022b). With a vapor pressure of 4.83 x 10"3 mm Hg at 20 °C (extrapolated), 8.3 x 10"
2 mm Hg at 25 °C (estimated), 8.4 mm Hg at 99.63 °C (measured), and a (measured) range of
4.80 x 10"3 mm Hg to 9.77 x 10"3 mm Hg, volatilization is not expected to be an important fate
process for this chemical (ATSDR, 2021, NCBI, 2022, EPA, 2022b). EPA's Toxics Release
Inventory reported release data for PFNA in 2020, with a total on-site disposal, off-site disposal,
and other releases concentration of 0 pounds from one facility (EPA, 2022c). PFNA is not listed
as a hazardous air pollutant (EPA, 2022d).
indoor Air
No studies from the U.S. reporting levels of PFNA in indoor air were identified from the primary
or gray literature.
Ambient Air
A single U.S. study measured levels of PFNA in ambient air (Kim and Kannan, 2007). Kim and
Kannan (2007) analyzed particle phase (n = 8) and gas phase (n = 8) concentrations of
perfluorinated acids in ambient air samples collected in and around Albany, New York in May
and July 2006 to examine the relative importance of certain media pathways to the contamination
of urban lakes. PFNA was detected in all gas phase samples with mean and median
concentrations of 0.21 pg/m3 and 0.20 pg/m3, respectively. PFNA was also detected in the
particulate phase, but the detection frequency was not reported. Authors reported particulate
phase mean and median concentrations of 0.13 pg/m3 and below the LOQ (0.12 pg/m3),
respectively.
Soil
The use and production of PFNA could result in its release to soils through various waste streams
(NCBI, 2022). When released to soil, PFNA is expected to have no mobility (NCBI, 2022).
PFNA has been measured in grass samples grown in soil containing PFNA and other PFAS near
Decatur, Alabama (ATSDR, 2021; Yoo et al., 2011). In addition, PFNA has been found to
accumulate in the roots of maize plants grown in soil containing PFNA and other PFAS
(Krippner et al., 2014; ATSDR, 2021).
Seven U.S. studies were identified that evaluated the occurrence of PFNA and other PFAS in soil
(Anderson et al., 2016; Blaine et al., 2013; Eberle et al., 2017; Galloway et al., 2020; Nickerson
et al., 2020; Venkatesan and Halden, 2014; Zhu and Kannan, 2019) (Table A-6). Among these
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studies, three analyzed soils potentially impacted by past AFFF use. The PFNA detection
frequencies varied widely (from less than 20% to over 90%) but mean concentrations tended to
be below 5 ng/g. Few studies analyzed soils in the vicinity of fluoropolymer manufacturing
facilities or by contaminated soil amendments. Other than control soils in two greenhouse and
field studies and one reference site, the U.S. studies did not evaluate soils without amendments
or without a nearby current or historical PFAS source.
Two studies analyzed soils in the vicinity of fluoropolymer manufacturing facilities (Galloway et
al., 2020; Zhu and Kannan, 2019). Galloway et al. (2020) collected soil samples in December
2016 and March 2018 near a fluoropolymer production facility outside Parkersburg, West
Virginia. The 2016 sampling included sites 4.0 km-48.1 km downwind to the north and
northeast of the facility and the 2018 sampling included sites 1.3 km-45.4 km north of the
facility. PFNA was detected in six of eight of the 2016 samples, however only one was above the
LOQ with a concentration of 1.63 ng/g. PFNA was also detected in six of seven of the 2018
samples, however only one was above the LOQ with a concentration of 1.92 ng/g at a distance of
1.3 km. Both the 2016 and 2018 samples that were above the LOQ were reported at the site
closest to the facility. In Zhu and Kannan (2019), authors studied PFAS concentrations in soil
contaminated by a nearby fluoropolymer manufacturing facility in Little Hocking, Ohio, that had
been manufacturing fluorochemicals for over five decades. The 45-acre well field located in a
floodplain meadowland was divided into quadrants and surface soil samples were collected from
multiple locations within each quadrant in October 2009. PFNA was detected in all 19 samples
with mean and median concentrations of 2.7 ng/g and 2.5 ng/g, respectively.
Three studies analyzed soils potentially impacted by AFFF use (Anderson et al., 2016; Eberle et
al., 2017; Nickerson et al., 2020). Anderson et al. (2016) assessed 40 sites across 10 active Air
Force installations throughout the continental United States and Alaska between March and
September 2014. Installations were included if there was known historic AFFF release in the
period 1970-1990. It is assumed that the measured PFAS profiles at these sites reflect the net
effect of several decades of all applicable environmental processes. The selected sites were not
related to former fire training areas and were characterized according to volume of AFFF
release—low, medium, and high. Across all sites, the PFNA detection frequency was 71.43% in
100 surface soil samples (median concentration of detects was 1.3 ng/g) and 14.42% in 112
subsurface soil samples (median concentration of detects was 1.5 ng/g). PFNA was detected
more frequently at high-volume release sites (50.8% in 32 surface soil samples with mean
concentration of 2.5 ng/g; 84.4% in 31 subsurface soil samples with mean concentration of
2.4 ng/g) than at low-volume sites (50.0% in 12 surface soil samples with mean concentration of
2.7 ng/g; 17.6% in 17 subsurface soil samples with mean concentration of 1.0 ng/g) and medium-
volume sites (38.3%) in 56 surface soil samples with mean concentration of 2.2 ng/g; 67.9% in 64
subsurface soil samples with mean concentration of 2.1 ng/g). Authors noted that given PFNA is
not present in 3M AFFF formulations, there may be some degree of telomer-based AFFF
contamination. Nickerson et al. (2020) developed a method to quantify anionic, cationic, and
zwitterionic PFAS from AFFF-impacted soils. The method was applied to two soil cores
collected from two different AFFF-impacted former fire training areas; the sampling year and
geographic location were not provided. Eleven soil samples, corresponding to 11 depths ranging
from 0.46 m to 15.1 m, were evaluated from Core E, and 12 soil samples, at depths ranging from
0.30 to 14.2 m, were evaluated from Core F. In Core E, PFNA was detected in 5 of 11 samples at
depths both at the surface and further below ground with PFNA concentrations ranging from
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below the LOQ to 1.96 ng/g dw. In Core F, PFNA was detected in 5 of 12 samples at the 5
depths closest to the surface, with concentrations ranging from below the LOQ to 4.17 ng/g dw
(LOQ not reported). Eberle et al. (2017) investigated the effects of an in situ chemical oxidation
treatment for remediation of chlorinated volatile organic compounds and PFAAs co-
contaminants. Soil samples were collected in 2012-2013 before and after a pilot scale field test
at a former fire training site at Joint Base Langley-Eustis, Virginia. Monthly fire training
activities were conducted at the site from 1968 to 1980 and irregular fire training activities
continued until 1990. Impacted soil was excavated in 1982 but details were not provided. PFNA
was detected in 1 of 5 pre-treatment samples and in 13 of 14 post-treatment samples. Of the
available three paired pre- and post-treatment soil samples, PFNA was not detected pre-treatment
in two pairings but detected post-treatment at 0.07 ng/g and 0.05 ng/g post-treatment. For the
third pairing, PFNA was detected at 1.1 ng/g pre-treatment and below the LOQ (0.06 ng/g) post-
treatment.
Of the remaining two studies conducted in the United States, Venkatesan and Halden (2014)
conducted outdoor mesocosm studies to examine the fate of PFAS in biosolids-amended soil
collected during 2005-2008. Biosolids were obtained from a wastewater treatment plant
(WWTP) in Baltimore that primarily treated wastewater from domestic sources with only minor
contribution (1.9%) from industry. The number of samples was not provided but PFNA was
detected in the control (nonamended) soil at levels below 0.5 ng/g dw and in the biosolids-
amended soil at a level not reported by the authors. In a field and greenhouse study, Blaine et al.
(2013) studied the uptake of PFAS into edible crops grown in control and biosolids-amended
soil. In the field study, urban biosolids were obtained from a WWTP receiving both domestic
and industrial waste while rural solids were obtained from a WWTP receiving domestic waste
only. PFNA was detected in soils from urban (mean = 0.20 ng/g, 0.28 n/g, and 0.40 ng/g in
control, 1 x14 and 2x amended fields, respectively) and rural fields (mean = 0.06 ng/g and
0.75 ng/g in control and 0.5 x amended fields, respectively). In the greenhouse study, three soils
(nonamended control, industrially impacted, and municipal) were investigated. Industrially
impacted soils contained composted biosolids from a small municipal WWTP that was impacted
by PFAA manufacturing while municipal soils were obtained from a reclamation site in Illinois
where municipal biosolids were applied for 20 years. PFNA was detected in all three soils at an
average concentration of 0.30 ng/g, 20.15 ng/g, and 6.11 ng/g in control, industrially impacted,
and municipal soil, respectively. Authors noted that the trace levels of PFAS detected in the
control soil may be due to minor cross-contamination from plowing, planting, or atmospheric
deposition from the surrounding area where biosolids have been applied.
14 0.5 x, lx; or 2x is defined as 'A I, or 2 times the agronomic rate of biosolids application to meet nitrogen
requirements of the crop.
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Table A-6. Summary of PFNA Data in Soil
Study
Location
Site Details
Results
United States
Galloway et al.
United States
Soil samples collected near a fluoropolymer
2016 sampling:
(2020)
(Parkersburg,
facility in two sampling trips in December 2016
Drag Strip Road (4.0 km) = 1.63 ng/g
West Virginia)
and March 2018. The 2016 sampling trip
Veto Lake (8.0 km) =
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Study
Location
Site Details
Results
United States
Anderson et al.
(2016)
United States Forty AFFF-impacted sites from ten active U.S.
(national) Air Force installations with historic AFFF
release between 1970 and 1990 that were not
related to former fire training areas. It is
assumed that the measured PFAS profiles at
these sites reflect the net effect of several
decades of all applicable environmental
processes. AFFF-impacted sites included
emergency response locations, hangers and
buildings, and testing and maintenance related
to regular maintenance and equipment
performance testing of emergency vehicles and
performance testing of AFFF solution. Previous
remedial activities for co-occurring
contaminants were not specifically controlled
for in the site selection process; active remedies
had not been applied at any of the sites selected.
Approximately ten samples were collected
between March and September 2014 at each site
for surface and subsurface soil; sites were
grouped according to volume of AFFF
release—low-volume typically had a single
AFFF release, medium-volume had one to five
releases, and high-volume had multiple
releases.
Surface soil:
Overall: n = 100, DF 71.43%, median
(maximum) =1.3 (23.0) ng/g
Breakdown by site:
Emergency Response (low-volume release):
n = 12, DF 50.0%, mean (range) = 2.7 (1.5-
4.1) ng/g
Hangars and Buildings (medium-volume release):
n = 56, DF 38.3%, mean (range) = 2.2 (0.21-
12) ng/g
Testing and Maintenance (high-volume release):
n = 32, DF 50.8%, mean (range) = 2.5 (0.24-
23) ng/g
(RL = 0.23 ng/g)
Subsurface soil:
Overall: n = 112, DF 14.42%, median
(maximum) =1.5 (6.49) ng/g
Breakdown by site:
Emergency Response (low-volume release):
n = 17, DF 17.6%, mean (range) = 1.0 (0.5-
1.5) ng/g
Hangars and Buildings (medium-volume release):
n = 64, DF 67.9%, mean (range) = 2.1 (0.21-
12) ng/g
Testing and Maintenance (high-volume release):
n = 31, DF 84.4%, mean (range) = 2.4 (0.24-
23) ng/g
(RL = 0.24 ng/g)
* Median calculated using quantified detections
*Non-detects were substituted with !/2 the reporting limit
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Study
Location
Site Details
Results
United States
Nickerson et al.
(2020)
United States Soil cores E and F from two different AFFF-
(unspecified) impacted fire training areas; sampling year and
geographic location not provided. Soil core E
contained 11- 0.3 m increment samples from
0.3-15.2 m below ground surface and was
collected in an area where the surficial soils
were likely disturbed due to regrading and other
soil redistribution activities. Soil core F
contained 12- 0.61 m increment samples from
0-14.2 m below ground surface and was
collected in an area where the surficial soils
were highly permeable only within the upper
0.5 to 1 m, and the underlying impermeable
clay layer exhibited a relatively high cation
exchange capacity and organic carbon content.
The water table was relatively shallow
(depth < 3 m) at both sites.
Core E:
0.46 m = 1.96 ng/g dw
2.9 m = < LOQ
3 .66 m = < LOQ
3.96 m = < LOQ
4.27 m = < LOQ
4.57 m = < LOQ
4.88 m = 0.22 ng/g dw
7.01 m = 0.26 ng/g dw
8.38 m = 0.73 ng/g dw
10.5 m = 1.09 ng/g dw
15 .1 m = < LOQ
Core F:
0.30 m = 0.70 ng/g dw
1.22 m = 4.17 ng/g dw
1.83 m = 3.23 ng/g dw
2.44 m = 1.04 ng/g dw
3.05 m = 0.64 ng/g dw
4.11 m = < LOQ
7.62 m = < LOQ
8.84 m = < LOQ
9.45 m = < LOQ
10.5 m = < LOQ
11.9 m = < LOQ
14.2 m = < LOQ
(LOQ not reported)
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Study
Location
Site Details
Results
United States
Eberle et al.
(2017)
United States
(Joint Base
Langley-Eustis,
Virginia)
Pilot testing area in former fire training area
(Training Site 15) at Joint Base Langley-Eustis
where monthly fire training activities were
conducted from 1968 to 1980 in a zigzag
pattern burn pit. Facility was abandoned in 1980
but irregular fire training activities using an
above-ground germed burn pit continued until
1990. Impacted soil was removed in 1982 but
additional details of the excavation are not well
known. Soil samples collected for pre- (April
and September 2012) and post- (December
2013) in situ chemical oxidation treatment using
a peroxone activated persulfate (OxyZone)
technology. Treatment was conducted in Test
Cell 1 over 113 days (April-August 2013). Soil
samples were collected adjacent to wells; wells
outside Test Cell 1 were used as sentry wells.
Well IDs for pre- and post-sampling were not
provided but the following three pairings were
assumed based on Table 2 in the paper: U-20
with SB-106; U-16 with SB-112; and 1-1 with
SB-109.
Pre-treatment:
1-1 (1.2-4.3 m) = 1.1 ng/g
1-2(1.2-4.3 m) =ND
U-12 (2.1 m) =ND
U-16 (3.0 m)=ND
U-20 (1.8 m)=ND
(LOQ = 0.68-0.72 ng/g)
Post-treatment:
SB-101 (4.3 m) = 0.07 ng/g
SB-105 (1.8 m) = 0.02 ng/g
SB-106/U-20 (1.8 m) = 0.07 ng/g
SB-106 (4.3 m) = 0.14 ng/g
SB-107 (1.8 m) = 0.03 ng/g
SB-107 (4.3 m) = 0.2 ng/g
SB-108 (1.8 m) = 0.03 ng/g
SB-108 (4.3 m) = 0.15 ng/g
SB-109/1-1 (3 m) =
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Study
Location
Site Details
Results
United States
Venkatesan and United States
Halden(2014) (Baltimore,
Maryland)
Archived agricultural soil (nonamended)
collected during 2005-2008 at a depth of 0-
20 cm from the United States Department of
Agriculture-Agricultural Research Service
Beltsville Agricultural Research Center; number
of sampling sites and number of samples not
provided.
Biosolids-amended soil obtained by mixing
biosolids and soil at a volumetric ratio of 1:2.
Biosolids were from Back River WWTP in
Baltimore, a full-scale activated sludge
treatment plant. Raw wastewater was primarily
from domestic sources with only minor
contribution (1.9%) from industry.
Nonamended: n = NR, DF NR, authors noted PFNA
concentration was between 0.1-0.5 ng/g dw
Amended: n = NR, DF NR, authors noted the detected
levels of PFNA, along with PFOA, PFNA, PFDA, and
PFUnA in the control soil accounted for 0.3-3% of
their initial levels in the amended soil mix
(MDL = 0.08 ng/g)
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Study
Location
Site Details
Results
United States
Blaine et al. United States Urban and rural full-scale field study with
(2013) (Midwest) control (nonamended) and biosolids-amended
plots. Three agricultural fields were amended
(0.5x, lx, or 2x) with municipal biosolids.
Urban biosolids (1 x and 2x) were from a
WWTP receiving both domestic and industrial
waste. Rural biosolids (0.5x) were from a
WWTP receiving domestic waste only. Control
plots were proximal to the rural and urban
amended corn grain and corn stover field sites;
sampling year not provided.
Greenhouse study with control (nonamended)
and biosolids-amended soil. Nonamended soil
obtained from a field that received commercial
fertilizers and had a similar cropping system as
the nearby municipal soil site. Municipal soil
was obtained from a reclamation site in Illinois
where municipal biosolids were applied at
reclamation rates for 20 years, reaching the
cumulative biosolids application rate of
1,654 Mg/ha. Industrially impacted soil was
created by mixing composted biosolids from a
small municipal (but impacted by PFAA
manufacturing) WWTP with control soil on a
10% mass basis. Sampling year not provided.
Field study:
Urban non-amended: n = 3-7, DF NR,
mean = 0.20 ng/g
Urban 1 x; n = 3-7, DF NR, mean = 0.28 ng/g
Urban 2x: n = 3-7, DF NR, mean = 0.40 ng/g
Rural non-amended: n = 3-7, DF NR, mean = .06 ng/g
Rural 0.5 x; n = 3-7, DF NR, mean = 0.75 ng/g
(LOQ not reported)
Greenhouse study:
Nonamended: n = 3-5, DF NR, mean = 0.30 ng/g
Industrially impacted: n = 3-5, DF NR,
mean = 20.15 ng/g
Municipal: n = 3-5, DF NR, mean = 6.11 ng/g
(LOQ not reported)
Notes: AFFF = aqueous film-forming foam; DF = detection frequency; dw = dry weight; LOQ = limit of quantitation; LHWA = Little Hocking Water Association; LOD = limit of
detection; MDL = method detection limit; 0.5X, 1 x; or 2x = !/2, 1, or 2 times the agronomic rate of biosolids application to meet nitrogen requirements of the crop; ND = not
detected; NR = not reported; RL = reporting limit; WWTP = wastewater treatment plant.
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Sediment
When released into water, PFNA is expected to adsorb to suspended solids and sediments
(NCBI, 2022). Concentrations of PFNA in sediment samples collected from the Hudson Bay
region of northeast Canada ranged from < 0.06 ng/g to 0.14 ng/g (dry weight) (Kelly et al., 2009;
NCBI, 2022).
Biomonitoring in the U.S. Population
CDC's NHANES results show that PFNA has been detected in > 95% of blood samples from
NHANES participants for most years evaluated (CDC, 2019, 2021 a,b, 2022). Whole-weight
serum levels of PFNA in the 50th percentile of the U.S. population for all years evaluated since
1999 were 0.600 |ig/L in 1999-2000 (detected in 96% of samples), 1.00 |ig/L in 2003-2004
(detected in 98.2% of samples), 1.10 |ig/L in 2005-2006 (detected in 99% of samples),
1.23 |ig/L in 2007-2008 (detected in 99.5% of samples), 1.23 |ig/L in 2009-2010 (detected in
99.8%) of samples), 0.860 |ig/L in 2011-2012 (detected in 99.2% of samples), 0.700 |ig/L in
2013-2014 (detected in 98.7% of samples), 0.600 |ig/L in 2015-2016 (detected in 98.7% of
samples), and 0.400 |ig/L in 2017-2018 (detected in 92% of samples) (CDC, 2019, 2021 a,b,
2022).
Recommended RSC
In summary, based on the physical properties, detected levels, and available exposure
information for PFNA, multiple non-drinking water sources (fish and shellfish, non-fish food,
some consumer products, indoor dust, and air) are potentially significant exposure sources.
Following the Exposure Decision Tree in EPA's 2000 Methodology (EPA, 2000), significant
potential sources other than drinking water ingestion were identified (Box 8A in the Decision
Tree); however, information is not available to quantitatively characterize exposure from these
different sources (Box 8B in the Decision Tree). Therefore, EPA recommends an RSC of 20%
(0.20) for PFNA.
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APPENDIX B. PFHxS: Summary of Occurrence in Water and Detailed
Relative Source Contribution.
Occurrence in Water
The production of PFHxS and its use as a raw material or precursor for manufacturing PFAS-
based products, as well as its previous use in firefighting foam and carpet treatment solutions,
could result in its release to the aquatic environment through various waste streams (NCBI,
2022). PFHxS has an estimated water solubility of 6,200 |ig/L (6.2 mg/L) at 25 °C and when
released to surface water, it is not expected to adsorb to suspended solids and sediment (NCBI,
2022). Volatilization from water surfaces is not expected to be an important fate process for
PFHxS (NCBI, 2022).
Drinking Water
Based on results from EPA's UCMR 3 monitoring, PFHxS has been detected in 1.12% of
drinking water systems in the United States, with mean and median concentrations of 140 ng/L
and 73 ng/L, respectively. The maximum concentration detected in drinking water systems was
1,600 ng/L (EPA, 2017). Drinking water samples collected from public drinking water systems
that are impacted by wastewater treatment effluent often contain higher concentrations of
perfluoroalkyls than samples collected from systems that are not impacted by wastewater
treatment effluent (Schultz et al., 2006a,b, as cited in ATSDR, 2021). For example, PFHxS was
detected in all samples collected from a public drinking water system in Los Angeles that was
highly impacted by wastewater treatment effluent; the mean concentrations of PFHxS in influent
and effluent samples were 5.1 ng/L (0.0051 |ig/L) and 6.1 ng/L (0.0061 |ig/L), respectively
(Quinones and Snyder, 2009 as cited in ATSDR, 2021). PFHxS has also been detected in the
municipal drinking water of communities located near a fluorochemical facility in Minnesota
(ATSDR, 2008). In comparison, no perfluoroalkyl chemicals were detected in influent or
finished drinking water samples collected from a public drinking water system in Aurora,
Colorado that was not highly impacted by wastewater treatment effluent (Quinones and Snyder,
2009 as cited in ATSDR, 2021).
For more information about PFHxS occurrence in drinking water, please see EPA (2022f).
A Standard of Quality for PFAS in bottled water is 0.005 |ig/L (5 ng/L) for one PFAS (e.g.,
PFHxS) and 0.0010 |ig/L (10 ng/L) for more than one PFAS (IBWA, 2022).
Groundwater
PFHxS was detected in each of the well water samples from a PFAS manufacturing facility in
Minnesota at concentrations ranging from 6,470 ng/L to 40,000 ng/L (6.47 |ig/L to 40.0 |ig/L)
(3M, 2007 as cited in ATSDR, 2021, NCBI, 2022). PFHxS was measured in offsite groundwater
near a PFAS manufacturing facility in Alabama at concentrations ranging from 12.7 ng/L to
622 ng/L (0.0127 |ig/L to 0.622 |ig/L) (3M, 2010 and Lindstrom et al., 2011 as cited in ATSDR,
2021). Median and maximum groundwater (i.e., not finished drinking water) concentrations of
870 ng/L and 290,000 ng/L (0.870 |ig/L and 290 |ig/L), respectively, were detected at 10 U.S.
military installations (Anderson et al., 2016 as cited in ATSDR, 2021).
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Surface Water
PFHxS in water bodies in New York was measured at concentrations ranging from 4.2 ng/L to
8.5 ng/L (0.0042 |ig/L to 0.0085 |ig/L) in Onondaga Lake (a Superfund site due to contamination
from industrial activity along its banks), from 2.5 ng/L to 5.6 ng/L (0.0025 |ig/L to 0.0056 |ig/L)
in Erie Canal, and from 0.9 ng/L to 2.8 ng/L (0.0009 |ig/L to 0.0028 |ig/L) in other lakes and
rivers (Sinclair et al., 2006 as cited in ATSDR, 2021 and NCBI, 2022). PFHxS concentrations
measured in lake water samples collected near Albany, New York ranged from < 0.25 ng/L to
4.05 ng/L (< 0.00025 |ig/L to 0.00405 |ig/L) (Kim and Kannan, 2007 as cited in ATSDR, 2021
and NCBI, 2022). PFHxS concentrations ranged from < 0.25 ng/L to 0.36 ng/L (< 0.00025 |ig/L
to 0.00036 |ig/L) in rain and snow samples collected in Albany, New York in 2006 to 2007 (Kim
and Kannan, 2007 as cited in NCBI, 2022). PFHxS was detected in more than 90% of 37 surface
water sites sampled across the northeastern United States in 2014 at a maximum concentration of
43 ng/L (0.043 |ig/L) at Mill Cove, Rhode Island, and a median concentration of 0.7 ng/L
(0.0007 |ig/L) (Zhang et al., 2016 as cited in ATSDR, 2021).
PFHxS concentrations in surface water collected from the Delaware River ranged from less than
the detection limit to 4.48 ng/L (0.00448 |ig/L) in 2007 to 2009 (DRBC, 2013 as cited in
ATSDR, 2021). PFHxS was measured in 100 samples collected from 80 sites from Cape Fear
Basin, North Carolina at an average concentration of 7.29 ng/L (0.00729 |ig/L), a median
concentration of 5.66 ng/L (0.00566 |ig/L), and a maximum concentration of 35.1 ng/L
(0.0351 |ig/L), with PFHxS not detected in 45.6% of the samples (Nakayama et al. 2007, as cited
in ATSDR, 2021).
PFHxS concentrations in water samples collected from Resolute Lake and Meretta Lake in 2003
and 2004 in the Canadian Arctic ranged from 1.5 ng/L to 24 ng/L (0.0015 |ig/L to 0.024 |ig/L)
(Stock et al., 2007 as cited in NCBI, 2022). PFHxS measured in surface water near a PFAS
manufacturing facility in Minnesota ranged from 93.6 ng/L to 4,580 ng/L (0.0936 |ig/L to
4.58 |ig/L) (3M, 2007 as cited in ATSDR, 2021). Median and maximum surface water
concentrations of 710 ng/L and 815,000 ng/L (0.710 |ig/L and 815 |ig/L), respectively, were
detected at 10 U.S. military installations (Anderson et al., 2016 as cited in ATSDR, 2021).
RSC for PFHxS
Literature Search and Screening
In 2020, EPA conducted a broad literature search to evaluate evidence for pathways of human
exposure to eight PFAS chemicals (PFOA, PFOS, PFBA, PFBS, PFDA, PFHxA, PFHxS, and
PFNA) (Holder et al., in prep). This search was not date limited and spanned the information
collected across the WOS, PubMed, and ToxNet/ToxLine (now ProQuest) databases. The results
of the PFHxS literature search of publicly available sources are available through EPA's Health
& Environmental Resource Online website at
https://hero.epa.eov/hero/index.cfm/proiect/paee/proiect id/2630.
The 950 literature search results for PFHxS were imported into SWIFT-Review (Sciome, LLC,
Research Triangle Park, NC) and filtered through the Evidence Stream tags to identify human
studies and non-human (i.e., those not identified as human) studies (Holder et al., in prep).
Studies identified as human studies were further categorized into seven major PFAS pathways
(Cleaning Products, Clothing, Environmental Media, Food Packaging, Home
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Products/Articles/Materials, Personal Care Products, and Specialty Products) as well as an
additional category for Human Exposure Measures. Non-human studies were grouped into the
same seven major PFAS pathway categories, except that the Environmental Media category did
not include soil, wastewater, or landfill. Only studies published between 2003 and 2020 were
considered. Application of the SWIFT-Review tags identified 654 peer-reviewed papers
matching these criteria for PFHxS.
Holder et al. (in prep) screened the 654 papers to identify studies reporting measured occurrence
of PFHxS in human matrices and media commonly related to human exposure (human
blood/serum/urine, drinking water, food, food contact materials, consumer products, indoor dust,
indoor and ambient air, and soil). For this synthesis, additional screening was conducted to
identify studies relevant to surface water (freshwater only) and groundwater using a keyword15
search for water terms.
Following the PECO criteria outlined in Table B-l, the title and abstract of each study were
independently screened for relevance by two screeners using litstream™. A study was included
as relevant if it was unclear from the title and abstract whether it met the inclusion criteria. When
two screeners did not agree if a study should be included or excluded, a third reviewer was
consulted to make a final decision. The title and abstract screening of Holder et al. (in prep) and
of this synthesis resulted in 494 unique studies being tagged as relevant (i.e., having data on
occurrence of PFHxS in exposure media of interest) that were further screened with full-text
review using the same inclusion criteria. After additional review of the evidence collected by
Holder et al. (in prep), 109 studies originally identified for other PFAS also contained
information relevant to PFHxS. Based on full-text review, 172 studies were identified as having
relevant, extractable data for PFHxS from the United States, Canada, or Europe for
environmental media, not including studies with only human biomonitoring data. Of these 172
studies, 161 were identified from Holder et al. (in prep), where primary data were extracted into
a comprehensive evidence database. Parameters of interest included: sampling dates and
locations, numbers of collection sites and participants, analytical methods, limits of detection and
detection frequencies, and occurrence statistics. Eleven of the 172 studies were identified in this
synthesis as containing primary data on only surface water and/or groundwater.
15 Keyword list: water, aquifer, direct water, freshwater, fresh water, groundwater, ground water, indirect water,
lake, meltwater, melt water, natural water, overland flow, recreation water, recreational water, river, riverine water,
riverwater, river water, springwater, spring water, stream, surface water, total water, water supply
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Table B-l. Populations, Exposures, Comparators, and Outcomes (PECO) Criteria
PECO Element Inclusion Criteria
Population Adults and/or children in the general population and populations in the
vicinity of PFAS point sources from the United States, Canada, or Europe
Exposure Primary data from peer-reviewed studies collected in any of the following
media: ambient air, consumer products, drinking water, dust, food, food
packaging, groundwater51, human blood/serum/urine, indoor air, landfill,
sediment, soil, surface watera (freshwater), wastewater/biosolids/sludge
Comparator Not applicable
Outcome Measured concentrations of PFHxS (or measured emissions from food
packaging and consumer products only)
Note:
a Surface water and groundwater were not included as relevant media in Holder et al. (in prep). Studies were re-screened for these
two media in this synthesis.
The evidence database of Holder et al. (in prep) additionally identified 18 studies for which the
main article was not available for review. As part of this synthesis, 17 of the 18 studies could be
retrieved. An additional three references were identified through gray literature sources,
described below, that were included to supplement the search results. The combined 20 studies
underwent full-text screening using the inclusion criteria in Table B-l. Based on full-text review,
five studies were identified as relevant.
Using the screening results from the evidence database and this synthesis, a total of 177 peer-
reviewed studies were identified as relevant. Fifty of these contained information relevant to the
U.S.
Additional Screening
EPA also searched the following gray literature sources for information related to relative
exposure of PFHxS for all potentially relevant routes of exposure (oral, inhalation, dermal) and
exposure pathways relevant to humans:
• AT SDR's Toxicological Profiles.,
• CDC's national reports on human exposures to environmental chemicals;
• EPA's CompTox Chemicals Dashboard;
• EPA's fish tissue studies;
• EPA's Toxics Release Inventory;
• EPA's UCMR data;
• Relevant documents submitted under the Toxic Substances Control Act and relevant
reports from EPA's Office of Chemical Safety and Pollution Prevention;
• U.S. Food and Drug Administration's (FDA's) Total Diet Studies and other similar
publications from FDA, U.S. Department of Agriculture, and Health Canada;
• NOAA's National Centers for Coastal Ocean Science data collections;
• National Science Foundation direct and indirect food and/or certified drinking water
additives;
• PubChem compound summaries;
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• Relevant sources identified in the relative source contribution discussions (Section 5) of
EPA's Proposed Approaches to the Derivation of a Draft Maximum Contaminant Level
Goal for Perfluorooctanoic Acid (PFOA)/Perfluorooctane Sulfonic Acid (PFOS) in
Drinking Water; and
• Additional sources, as needed.
EPA has included available information from these gray literature sources for PFHxS relevant to
its uses, chemical and physical properties, and for occurrence in drinking water (directly or
indirectly in beverages like coffee, tea, commercial beverages, or soup), ambient air, foods
(including fish and shellfish), incidental soil/dust ingestion, and consumer products. EPA has
included available information specific to PFHxS below on any regulations that may restrict
PFHxS levels in media (e.g., water quality standards, air quality standards, food tolerance
levels).
Summary of Potential Sources of PFHxS Exposure
EPA presents information below from studies performed in the United States. While studies from
non-U. S. countries inform an understanding global exposure sources and trends, the RSC
determination is based on the available data for the United States.
Dietary Sources
Seafood
PFHxS was detected in 71 of 157 fish tissue composite samples collected during EPA's National
Lake Fish Tissue Study, with a maximum concentration of 3.50 ng/g and a 50th percentile
concentration of < 0.12 ng/g (Stahl et al., 2014). It was not detected in the 162 fish tissue
composite samples collected during EPA's 2008-2009 National Rivers and Streams Assessment
(NRSA) (Stahl et al., 2014). More recently, PFHxS was detected in 32 of 349 fish tissue
composite samples at concentrations ranging from 0.121 ng/g to 0.980 ng/g in EPA's 2013-2014
NRSA (EPA, 2020). PFHxS was also detected in 1 of 152 fish tissue composite samples at a
concentration of 0.96 ng/g in EPA's 2015 Great Lakes Human Health Fish Fillet Tissue Study
(EPA, 2021). PFHxS has been detected in a mixture of fish fillet samples collected from
Mississippi River sites in Minnesota at a concentration of 0.47 ng/g (Delinsky et al., 2010;
ATSDR, 2021). PFHxS has been detected in Irish pompano (Diapterus auratus), silver porgy
(Diplodus argenteus), and grey snapper (Lutjanus griseus) from the St. Lucie Estuary in in
NOAA's National Centers for Coastal Ocean Science, National Status and Trends Data (NOAA,
2022).
Five additional U.S. studies were identified that evaluated PFHxS levels in seafood (Byrne et al.,
2017; Chiesa et al., 2019; Schecter et al., 2010; Young et al., 2013, 2022) (Table B-2). One study
evaluated fish samples collected directly from rivers and lakes (Byrne et al., 2017). As part of a
study to assess exposure to PFHxS and other PFAS among residents of two remote Alaska
Native villages on St. Lawrence Island, Byrne et al. (2017) measured PFAS concentrations in
stickleback and Alaska blackfish, resident fish used as sentinel species to detect accumulation of
PFAS in the local environment. Stickleback were collected from three locations—Suqitughneq
(Suqi) River watershed (n = 9 composite samples), Tapisaggak (Tapi) River (n = 2 composite
samples), and Troutman Lake (n = 3 composite samples). Blackfish were collected from the Suqi
River (n = 29) but were not found in the other water bodies. Authors reported that the Suqi River
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watershed was upstream and downstream of a formerly used defense site and Tapi River was
approximately 5 km east of a military site, however at the start of the study none of the sites were
known to be contaminated with PFAS. The sample dates were not reported. PFHxS was not
detected in any of the stickleback and blackfish samples, despite the authors noting that
stickleback from Troutman Lake had "exceptionally high" total PFAS concentrations.
The remaining four studies purchased seafood from stores and fish markets (Chiesa et al., 2019;
Schecter et al., 2010; Young et al., 2013, 2022). Young et al. (2013) assessed fish and shellfish
collected in 2010-2012 from retail markets across the continental United States. Retail markets
in California, Florida, Illinois, Mississippi, New Jersey, New York, Tennessee, Texas, and
Washington, D.C. were represented. Authors selected the ten most consumed fish and shellfish
in the United States that were farm raised, wild caught, or had unknown origin. Among the crab
meat, shrimp, striped bass, catfish, clams, cod, flounder, pangasius, pollock, tuna, salmon,
scallops, and tilapia, PFHxS was only detected in one of ten samples of striped bass at a
concentration of 0.66 ng/g. Young et al. (2022) evaluated fish and shellfish purchased from retail
markets in the Washington. D.C. metropolitan area and online markets (clams only) from March
2021 through May 2022. Seafood samples represented eight of the top ten consumed fish and
shellfish in the United States. Seafood samples were farm raised, wild-caught or of unknown
origin, and location of harvest was provided when known. PFHxS was only detected in two
seafood types, crab and clam meat. All samples of clam meat (n=10) had detectable
concentrations of PFHxS, ranging from 51-605 ng/kg. Only two samples of crabs (n=l 1) had
detectable levels of 112 and 242 ng/kg. Authors also analyzed food packaging for PFAS analytes
and did not identify any packaging samples with detectable levels of PFAS. Schecter et al.
(2010) evaluated PFHxS and other PFAS in seafood collected from five Dallas, Texas grocery
stores in 2009. The origin or source of seafood was not described. Seafood included canned
sardines in water, canned tuna, fresh catfish fillet, cod, frozen fish sticks, salmon, and tilapia
(n = 1 composite sample for each seafood type). PFHxS was only detected in cod at a
concentration of 0.07 ng/g ww. Finally, in a multicontinental study, Chiesa et al. (2019) collected
salmon from a wholesale fish market in Milan, Italy; the sampling year was not reported. Wild-
caught salmon samples originated from the United States (n = 7), Canada (n = 15), and Scotland
(n = 2), while farmed salmon samples originated from Norway (n = 25) and Scotland (n = 17).
Among the salmon that originated from the United States Pacific Ocean (FAO 67 and 77), two
species—Oncorhynchus kisutch and Oncorhynchus keta—were analyzed, with PFHxS not
detected in either species (LOQ = 0.015 ng/g). PFHxS was also not detected in wild-caught
salmon from Canada and Scotland.
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Table B-2. Summary of PFHxS Data in Seafood
Study
Location and Source
Seafood Type
Results
United States
Byrne et al. (2017)
United States (Alaska)
Stickleback collected from three
locations on St. Lawrence Island:
Suqitughneq (Suqi) River watershed
(upstream and downstream of a formerly
used defense site), Tapisaggak (Tapi)
River (located approximately 5 km east
of military site), and Troutman Lake, a
coastal lake situated adjacent to the
village of Gambell.
Alaska blackfish collected from the Suqi
River but were absent from the other
water bodies.
Sampling year not reported. No sites
were known to be contaminated with
PFASs at the initiation of the study.
Stickleback and Alaska
blackfish
Stickleback:
Troutman Lake: n = 3*; DF 0%
Suqi River: n = 9*; DF 0%
Tapi River: n = 2*; DF 0%
Blackfish: n = 29; DF 0%
(LOQ = 0.5-1 ng/g ww for all PFAS)
*Number of composite samples, each composed of
-10 stickleback fish
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Study
Location and Source
Seafood Type
Results
United States
Young et al. (2013)
United States (California; Illinois;
Mississippi; Tennessee; Florida; New
Jersey; New York; Texas; Washington,
DC.)
Fish and shellfish collected from retail
markets in 11 areas across the
continental United States from 2010-
2012. The fish and shellfish included
farm raised, wild caught, and unknown
origin, as well as freshwater fish,
saltwater fish, and euryhaline fish.
Crab meat, clams, cod, flounder,
pangasius, salmon, scallops, and tilapia
purchased from Washington, D.C.
Shrimp purchased from Orlando,
Florida; Memphis, Tennessee; and
Nashville, Tennessee. Striped bass
purchased from New York, New York
and Cherry Hill, New Jersey. Catfish
purchased from Indianola, Mississippi;
Dallas, Texas; Tampa, Florida; and
Orlando, Florida. Pollock purchased
from Huntington Beach, California.
Tuna purchased from Chicago, Illinois.
Crab, shrimp, striped bass,
catfish, clams, cod,
flounder, pangasius,
pollock, tuna (can and
pouch), salmon, scallops
(bay and sea), tilapia
Striped bass: n = 10, DFa 10%, range = ND-
0.66* ng/g
Crab meat: n = 1, DF 0%
Shrimp: n = 9, DF 0%
Catfish: n = 13, DF 0%
Clams: n = 1, DF 0%
Cod: n = 1, DF 0%
Flounder: n = 1, DF 0%
Pangasius: n = 1, DF 0%
Pollock: n = 1, DF 0%
Tuna: n = 3, DF 0%
Salmon: n = 2, DF 0%
Scallops: n = 2, DF 0%
Tilapia: n = 1, DF 0%
(MDL = 0.55 ng/g for all seafood)
*This value was above the MDL but below the
LOQ; LOQ is estimated as 3x the MDL
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Study
Location and Source
Seafood Type
Results
United States
Young et al. (2022)
United States (Washington, D.C.)
Fish and shellfish collected from retail
markets in the Washington, D.C.
metropolitan area from 2021-2022. Fish
and shellfish samples included farm
raised, wild-caught and unknown origin.
Country of origin was provided, if
known.
Ten samples of each seafood type,
except for crab, which included 11
samples.
Crab, clams (can), shrimp,
cod, pollock (fish sticks,
fillet), salmon, tuna (can
and pouch), tilapia
Clams: n=10, DF 100%, range=51-605 ng/kg
Crab meat: n=ll, DF 20%, range=
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Study Location and Source Seafood Type Results
United States
Chiesa et al. (2019) United States (Pacific Ocean) Wild-caught salmon Oncorhynchus kisutch: n = 5, DF 0%
Wild-caught fish were collected at a (Oncorhynchus kisutch Oncorhynchus keta: n = 2, DF 0%
wholesale fish market in Milan, Italy. an<^ Oncorhynchus keta) (LOQ = 0.015 ng/g)
Sampling year was not reported. The
wild-caught salmon were from USA-
Pacific Ocean (Food and Agriculture
Organization Area 67 and 77).
Notes: DF = detection frequency; LOD = limit of detection; LOQ = limit of quantitation; MDL = method detection limit; ND = not detected; NR = not reported; ww = wet weight.
Bold indicates detected levels of PFHxS in food.
a The DF and/or mean was not reported in the study and was calculated in this synthesis. Means were calculated only when DF = 100%.
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Other Food Sources
PFHxS was included in a suite of PFAS evaluated in FDA's 2019, 2021, and 2022 Total Diet
Study Sampling (U.S. FDA, 2020a,b, 2021a,b, 2022a,b); however, it was not detected in any of
the food samples tested. It should be noted that FDA indicated that the sample sizes used in the
PFAS 2019, 2021, and 2022 Total Diet Study Sampling were limited and that the results should
not be used to draw definitive conclusions about PFAS levels in the general food supply (U.S.
FDA, 2022c). PFHxS was detected in milk samples collected from a farm with groundwater
known to be contaminated with PFAS; however, it was not detected in produce collected from an
area near a PFAS production plant, in FDA studies of the potential exposure to the U.S.
population to PFAS (U.S. FDA, 2018, 2021c). PFHxS is not a registered pesticide under the
Federal Insecticide, Fungicide, and Rodenticide Act (FIFRA), and EPA does not set a 40 CFR
Part 180 pesticide tolerance in food and feed commodities for PFHxS (U.S. GPO, 2022).
Maximum residue levels for PFHxS were not found in the Global Maximum Residue Level
Database (Bryant Christie Inc., 2022).
Nine peer-reviewed studies were identified that examined PFHxS in food sources other than
seafood, with 2 in breastmilk and 7 in food types other than breastmilk (Blaine et al., 2013, 2014;
Genualdi et al., 2017; Kuklenyik et al., 2004; Schecter et al., 2010; Scher et al., 2018; Tipton et
al., 2017; von Ehrenstein et al., 2009; Young et al., 2012) (Table B-3). Few U.S. studies
analyzed foods from any one origin—four sampled crops grown in areas with known or
suspected PFAS contamination, including biosolids-amended soils, two sampled from crops as
part of greenhouse and field studies, one studied wild-caught alligator meat. Only two studies
sampled from store- or market-bought meats, eggs, produce, and dairy.
Scher et al. (2018) evaluated garden produce samples from homes in Minnesota within and
outside of a GCA in the vicinity of a former 3M PFAS production facility. Twenty homes within
the GCA had previous or ongoing PFAS contamination in drinking water and were served by the
Oakdale, Minnesota public water system or a private well previously tested and shown to have
detectable levels of PFOA or PFOS. A total of 279 produce samples (232 inside GCA, 47 outside
GCA) were collected between May and October 2010. PFHxS was detected in 1% of the 232
produce samples from inside the GCA (1 floret sample and 1 leaf sample). The authors suggested
that the two detections were associated with PFAS present in irrigation water that had
accumulated in produce. They also noted that accumulation of PFAS was particularly high in
florets. Three homes that were outside the GCA served as a reference. No PFHxS was detected
in produce samples from home gardens outside the GCA. Genualdi et al. (2017) analyzed PFAS
contamination in a Massachusetts cranberry bog approximately 10 miles from a military base
with a history of AFFF usage. Ten cranberry samples were taken directly from trucks
transporting cranberries and 32 cranberry samples were collected directly from the bog water in
November 2016. PFHxS was not detected in any samples (MDL = 0.79 ng/g).
Two studies purchased food items from stores and markets for evaluation (Schecter et al., 2010;
Young et al., 2012). Schecter et al. (2010) assessed PFHxS and other PFAS in food samples
collected from five Dallas, Texas grocery stores in 2009. The origin or source of each food was
not described. Food items included meat products (bacon, canned chili, chicken breast, ground
beef, roast beef, ham, sausage, and turkey), dairy (butter, cheeses, frozen yogurt, ice cream, milk,
and yogurt), eggs, grains (cereal), fruits and vegetables (apples, potatoes), and fats/other (canola
oil, margarine, olive oil, peanut butter). PFHxS was not detected in any of the food samples. In
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Young et al. (2012), cow milk was purchased from retail markets across the continental United
States representing 17 states; the sampling year was not reported. Cow milk samples included
organic milk, vitamin D added milk, and ultra-pasteurized milk. PFHxS was not detected in any
of the 49 retail milk samples (MDL = 0.15 ng/g).
One study investigated PFAS levels from wild meat (Tipton et al., 2017). Tipton et al. (2017)
assessed alligator tail meat that was collected during the South Carolina recreational hunting
season between September to October 2015. Tail meat samples were collected from four
different public hunt units—Southern Coastal, Middle Coastal, Midlands, and Pee Dee. PFHxS
was detected in all samples from all hunt units. Median concentrations from Southern Coastal
(n = 19), Middle Coastal (n = 17), Midlands (n = 5), and Pee Dee (n = 2) were 0.087 ng/g,
0.099 ng/g, 0.0816 ng/g, and 0.093 ng/g wet mass, respectively.
Two studies by Blaine et al. (2013, 2014) evaluated PFHxS in crops grown in greenhouse and
field studies. In Blaine et al. (2014), PFAS levels were measured in celery root, pea fruit, and
radish root grown in a greenhouse with control (unamended) soil, industrially impacted soil, and
municipal soil (n = 3-5). PFHxS was detected in radish root from all three soils, celery shoot
from the industrially impacted and municipal soil, and pea fruit from only industrially impacted
soil. Mean concentrations of PFHxS in radish root for the control, industrially impacted, and
municipal soil were 3.81 ng/g, 2.84 ng/g, and 4.33 ng/g, respectively. Mean concentrations of
PFHxS in celery shoot for the industrially impacted and municipal soil were 3.19 ng/g and
0.38 ng/g, respectively. The mean concentration of PFHxS in pea fruit in the industrially
impacted soil was 0.24 ng/g. Authors noted minor cross-contamination of the control soil due to
the proximity of the unamended soil to biosolids-amended soil. In Blaine et al. (2013), authors
studied the uptake of PFAS into edible crops in both field and greenhouse studies. In the field
study, PFAS levels were measured in corn grain and corn stover grown with control
(unamended), urban biosolids-amended, and rural biosolids-amended soil (n = 3-7). Mean
PFHxS concentrations were below the LOQ in both corn grain and corn stover grown in any
field study plots (< 0.04 ng/g for corn grain; < 0.29 ng/g for corn stover). In the greenhouse
study, lettuce and tomato plants were grown in control soil, industrially impacted soil, or
municipal soil (n = 3-5). Mean PFHxS concentrations were below the LOQ for lettuce and
tomato grown in the control soil and for tomato grown in municipal soil; however, mean PFHxS
levels were 10.44 ng/g and 5.54 ng/g for lettuce grown in industrially impacted and municipal
soils, respectively, and 0.76 ng/g for tomato grown in industrially impacted soil. Sampling year
was not reported.
The remaining two studies evaluated the occurrence of PFHxS in breastmilk (Kuklenyik et al.,
2004; von Ehrenstein et al., 2009). von Ehrenstein et al. (2009) collected breastmilk samples
between December 2004 and July 2005 from women between the ages of 18 and 38 at the time
of recruitment as part of the pilot study Methods Advancement for Milk Analysis (MAMA).
Women provided milk samples at two visits—the first visit was 2-7 weeks postpartum, and the
second visit was 3-4 months postpartum. PFHxS was not detected in any of the samples from the
first visit (n = 18) or second visit (n = 20). Similarly, PFHxS was below the LOD (0.3 ng/mL) in
the samples reported by Kuklenyik et al. (2004). Kuklenyik et al. (2004) did not report
information on the breastmilk donors or the sampling procedure as it was unavailable; PFHxS
was not detected in either of the two samples.
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Table B-3. Summary of PFHxS Data in Other Food
Study
Location and Source
Food Types
Results
United States
Scher et al. (2018) United States (Minnesota)
Home garden produce samples were
collected between May and October 2010
from 20 homes in 3 cities within a GCA as
well as 3 homes in the Twin Cities Metro
outside the GCA. Homes within the GCA
were near a former 3M PFAS production
facility, had previous or ongoing PFAS
contamination in drinking water, and were
served by the Oakdale, Minnesota public
water system or were formerly or currently
using a private well previously tested and
shown to have detectable levels of PFOA
or PFOS.
279 produce samples (232 within GCA
and 47 outside GCA) consisting of mature,
edible portions of plants were analyzed.
Plant part categories included floret, fruit,
leaf, root, seed, and stem.
Fruits and
vegetables
Within GCA:
All: n = 232, DF 1%, median (range) = ND (ND-
0.066) ng/g
Floret: n = 5, DF 20%, median (range) = ND (ND-
0.066) ng/g
Leaf: n = 35, DF 3%, median (range) = ND (ND-
0.046) ng/g
Garden fruit (n = 98), yard fruit (n = 13), root (n = 29),
seed (n = 29), and stem (n = 23): DF 0%
Outside GCA:
All: n = 47, DF 0%
Floret (n = 1), garden fruit (n = 15), yard fruit (n = 4),
leaf (n = 12), root (n = 5), seed (n = 5), and stem
(n = 5): DF 0%
(MDL = 0.003 to 0.029 ng/g depending on the analyte and
type of produce)
Genualdi et al. United States (Massachusetts)
(2017) Samples from cranberry bog with surface
water contaminated with PFAS—likely
due to proximity to a military base with a
history of AFFF usage. The bog was
located approximately 10 miles from the
military base. Ten cranberry samples taken
directly from trucks transporting
cranberries (five samples each from two
trucks) and 32 cranberry samples taken
directly from 12 sections of the bog water.
Samples collected in November 2016.
Fruits
n = 42, DF 0%
(MDL = 0.79 ng/g)
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Study
Location and Source
Food Types
Results
United States
Schecter et al. United States (Texas)
(2010) Food samples from five different grocery
stores in Dallas, Texas were collected in
2009. Ten individual samples were
collected for each food type and combined
to form composite samples. The
origin/source of the food samples were not
reported.
Dairy; fruits and
vegetables; grains;
meat; fats/other
Meat
Hamburger: n = 1, DF 0%, LOD = 0.04 ng/g ww
Bacon: n = 1, DF 0%, LOD = 0.05 ng/g ww
Sliced turkey: n = 1, DF 0%, LOD = 0.02 ng/g ww
Sausages: n = 1, DF 0%, LOD = 0.04 ng/g ww
Ham: n = 1, DF 0%, LOD = 0.02 ng/g ww
Sliced chicken breast: n = 1, DF 0%, LOD = 0.02 ng/g
ww
Roast beef: n = 1, DF 0%, LOD = 0.02 ng/g ww
Canned chili: n = 1, DF 0%, LOD = 0.01 ng/g ww
Dairy and Eggs
Butter: n = 1, DF 0%, LOD = 0.09 ng/g ww
American cheese: n = 1, DF 0%, LOD = 0.04 ng/g ww
Other cheese: n = 1, DF 0%, LOD = 0.04 ng/g ww
Whole milk: n = 1, DF 0%, LOD = 0.02 ng/g ww
Ice cream: n = 1, DF 0%, LOD = 0.03 ng/g ww
Frozen yogurt: n = 1, DF 0%, LOD = 0.02 ng/g ww
Whole milk yogurt: n = 1, DF 0%, LOD = 0.03 ng/g ww
Cream cheese: n = 1, DF 0%, LOD = 0.02 ng/g ww
Eggs: n = 1, DF 0%, LOD = 0.04 ng/g ww
Grains
Cereals: n = 1, DF 0%, LOD = 0.04 ng/g ww
Fruits and Vegetables
Apples: n = 1, DF 0%, LOD = 0 .02 ng/g ww
Potatoes: n = 1, DF 0%, LOD = 0.04 ng/g ww
Fats/Other
Olive oil: n = 1, DF 0%, LOD = 0.3 ng/g ww
Canola oil: n = 1, DF 0%, LOD = 0.5 ng/g ww
Margarine: n = 1, DF 0%, LOD = 0.03 ng/g ww
Peanut butter: n = 1, DF 0%, LOD = 0.03 ng/g ww
*n reflects number of composite samples, each composed
of -10 individual samples
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Study
Location and Source
Food Types
Results
United States
Young et al.
United States (17 states)
Dairy
n = 49, DF 0%,
(2012)
Retail cow's milk samples were all
pasteurized whole milk, commercially
available, and purchased at retail markets
across the continental United States
representing 17 states. Samples were
organic milk, vitamin D added milk, and
ultra-pasteurized milk. Sampling year not
reported.
(MDL = 0.15 ng/g)
Tipton et al.
United States (South Carolina)
Meat
Alligator tail:
(2017)
Alligator tail meat samples were collected
from a local wild game meat processer
during the South Carolina recreational
hunt season between September to October
2015. Samples were from four different
public hunt units—Southern Coastal,
Middle Coast, Midlands, and Pee Dee.
Southern coastal: n = 19, DFa 100%, median
(range) = 0.087 (0.051-0.252) ng/g wet mass
Middle coastal: n = 17, DFa 100%, median
(range) = 0.099 (0.063-0.272) ng/g wet mass
Midlands: n = 5, DFa 100%, median (range) = 0.0816
(0.054-0.158) ng/g wet mass
Pee Dee: n = 2, DFa 100%, median (range) = 0.093
(0.071-0.115) ng/g wet mass
(RL not reported)
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Study
Location and Source
Food Types
Results
United States
Blaine et al. United States (Midwest)
(2014) Crops grown in in greenhouse study with
control (unamended), industrially impacted
soil, or municipal soil. Control soil had
minor cross-contamination due to
proximity to biosolids-amended fields.
Industrially impacted soil was amended
with industrially impacted biosolids, and
municipal soil was amended with
municipal biosolids for over 20 years.
Crops grown in the greenhouse study were
grown from seed in pots, which were
randomly arranged within the greenhouse.
Sampling year not reported.
Fruits and
vegetables
Radish root:
Control: n = 3-5, DF NR, mean = 3.81 ng/g
Industrially impacted; n = 3-5, DF NR,
mean = 2.84 ng/g
Municipal: n = 3-5, DF NR, mean = 4.33 ng/g
Celery shoot:
Control: n = 3-5, DF 0%
Industrially impacted: n = 3-5, DF NR,
mean = 3.19 ng/g
Municipal: n = 3-5, DF NR, mean = 0.38 ng/g
Pea fruit:
Control: n = 3-5, DF 0%
Industrially impacted: n = 3-5, DF NR,
mean = 0.24 ng/g
Municipal: n = 3-5, DF 0%
(LOQ = 0.03 ng/g)
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Study
Location and Source
Food Types
Results
United States
Blaine et al. United States (Midwest)
(2013) Crops grown in urban and rural full-scale
field study with control (unamended) and
biosolids-amended soil. Three agricultural
fields were amended (0.5x, lx, or 2x) with
municipal biosolids. Urban biosolids (lx
and 2x) were from a WWTP receiving
both domestic and industrial waste. Rural
biosolids (0.5 x) were from a WWTP
receiving domestic waste only. Control
plots were proximal to the rural and urban
amended corn grain and corn stover field
sites; sampling year not provided.
Crops grown in greenhouse study with
control (nonamended) and biosolids-
amended soil. Nonamended soil obtained
from a field that received commercial
fertilizers and had a similar cropping
system as the nearby municipal soil site.
Municipal soil was obtained from a
reclamation site in Illinois where
municipal biosolids were applied at
reclamation rates for 20 years, reaching the
cumulative biosolids application rate of
1,654 Mg/ha. Industrially impacted soil
was created by mixing composted
biosolids from a small municipal (but
impacted by PFAA manufacturing)
WWTP with control soil on a 10% mass
basis. Sampling year not provided.
Fruits and
vegetables; grains
Field study:
Corn grain:
Urban nonamended: n = 3-7, DF NR,
mean = < 0.04 ng/g
Urban 1 x; n = 3-7, DF NR, mean = < 0.04 ng/g
Urban 2x: n = 3-7, DF NR, mean = < 0.04 ng/g
Rural nonamended: n = 3-7, DF NR,
mean = < 0.04 ng/g
Rural 0.5 x; n = 3-7, DF NR, mean = < 0.04 ng/g
Corn stover:
Urban nonamended: n = 3-7, DF NR,
mean = < 0.29 ng/g
Urban 1 x; n = 3-7, DF NR, mean = < 0.29 ng/g
Urban 2x: n = 3-7, DF NR, mean = < 0.29 ng/g
Rural nonamended: n = 3-7, DF NR,
mean = < 0.29 ng/g
Rural 0.5 x: n = 3-7, DF NR, mean = < 0.29 ng/g
(LOQ = 0.04 ng/g for corn grain; LOQ = 0.29 ng/g for
corn stover)
Greenhouse study:
Lettuce:
Nonamended: n = 3-5, DF NR, mean = <0.01 ng/g
Industrially impacted: n = 3-5, DF NR,
mean = 10.44 ng/g
Municipal: n = 3-5, DF NR, mean = 5.54 ng/g
Tomato:
Nonamended: n = 3-5, DF NR, mean = < 0.03 ng/g
Industrially impacted: n = 3-5, DF NR,
mean = 0.76 ng/g
Municipal: n = 3-5, DF NR, mean = < 0.03 ng/g
(LOQ = 0.01 ng/g for lettuce; LOQ = 0.03 ng/g for
tomato)
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Study
Location and Source
Food Types
Results
United States
von Ehrenstein et
United States (North Carolina)
Breastmilk
Visit #1: n= 18, DF 0%
al. (2009)
As part of the Methods Advancement for
Milk Analysis (MAMA) pilot study, 34
breastfeeding women aged 18 to 38 years
at recruitment provided breastmilk samples
at two visits. The first visit occurred 2-
7 weeks postpartum, and the second visit
occurred 3-4 months postpartum. Both
visits were between December 2004 and
July 2005.
Visit #2: n = 20, DF 0%
(LOQ = 0.30 ng/mL)
Kuklenyik et al.
United States (Georgia)
Breastmilk
n = 2, DF 0%
(2004)
Authors reported that no information was
provided on the human milk donors or the
sampling procedure.
(LOD = 0.3 ng/mL)
Notes: AFFF = aqueous film-forming foam;DF = detection frequency; GCA = groundwater contamination area; LOD = limit of detection; LOQ = limit of quantitation;
MDL = method detection limit; ND = not detected; NR = not reported; RL = reporting limit; ww = wet weight; WWTP = wastewater treatment plant.
Bold indicates detected levels of PFHxS in food.
a The DF and/or mean was not reported in the study and was calculated in this synthesis. Means were calculated only when DF = 100%.
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Consumer Products
PFHxS has been used in laboratory applications and as a raw material or a precursor for the
manufacture of PFAS/perfluoroalkyl sulfonate-based products, though production of PFHxS in
the United States was phased out by its major manufacturer in 2002 (Backe et al., 2013, Buck et
al., 2011; OECD, 2011 and Sigma-Aldrich, 2014; NCBI, 2022). PFHxS has also been used in
firefighting foam and carpet treatment solutions, and it has been used as a stain and water
repellant (Garcia and Harbison, 2015; NCBI, 2022). PFHxS has been detected in aqueous film
forming foam, aftermarket carpet protection products, chipboards, leather, membranes for
apparel, treated apparel, and photoprint ink and laser ink (Backe et al., 2013; Becanova, et al.,
2016; Buck et al., 2011; Gliige et al., 2021; Herzke et al., 2009; Kotthoff et al., 2015; Liu et al.,
2014; NCBI, 2022; Norwegian Environment Agency, 2018).
Two studies were identified that analyzed PFHxS concentrations in a range of consumer
products, including children's nap mats, household carpet/fabric-care liquids, and textiles (Liu et
al., 2014; Zheng et al., 2020) (Table B-4). Few U.S. studies analyzed children's products, fabric
treatments, treated fabrics, sealants, and similar products, and none of the U.S. studies reviewed
sampled for PFAS in other household products and articles such as cosmetics, cleaners, paints,
upholstered furniture, etc. Of the U.S. studies, the majority of the consumer products evaluated
are likely used by adults (e.g., floor waxes), can come into contact with both adults and children
(e.g., treated upholstery), or the user was not specified (e.g., clothing).
Zheng et al. (2020) determined the occurrence of ionic and neutral PFAS in the childcare
environment (dust and nap mats). Samples of children's nap mats were collected from seven
Seattle childcare centers (n = 26; 20 polyurethane foam, 6 vinyl cover samples). PFHxS was
detected in 73% of nap mat samples with a mean concentration of 0.32 ng/g. Half of the
analyzed mats were purchased as new products and the other half were used. The authors
reported that total PFAS levels in the new vs. used mats were not significantly different. Total
PFAS levels in mat foam vs. mat covers were also similar. Based on these results, the authors
suggested that indoor air was not the major source of PFAS in mats and that PFAS in mats could
be associated with the manufacturing process.
Liu et al. (2014) analyzed the occurrence of PFAS in consumer products (including commercial
carpet/fabric-care liquids, household carpet/fabric-care liquids, treated apparel, treated home
textile and upholstery, treated floor waxes and stone-wood sealants, membranes for apparel, and
thread-sealant tapes and pastes) purchased between March 2007 and September 2011 from local
retailers and online stores in the United States. PFHxS was analyzed in a subset of these
consumer products, originating from the United States, England, Dominican Republic, Vietnam,
and China, and was detected in one out of two commercial carpet/fabric-care liquids samples at
194 ng/g, in two out of four household carpet/fabric-care liquids and foams samples at 88.8 ng/g
and 155 ng/g, in one out of two treated children's apparel samples at 1.70 ng/g (in boy's uniform
pants), in one out of two treated home textile and upholstery samples at 12.1 ng/g, in one apparel
membrane sample at 7.10 ng/g, and in one out of two thread-sealant tapes and pastes samples at
60.3 ng/g. PFHxS was not detected in one treated floor wax and stone/wood sealant sample.
Detection limits were not reported in the study.
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Table B-4. Summary of PFHxS Consumer Product Data
Study
Location
Site Details
Results
United States
Zheng et al. (2020)
United States (Seattle,
Washington)
Children's nap mat samples (n = 26, finely cut)
from seven Seattle childcare centers, including
polyurethane foam (n = 20) and vinyl cover
(n = 6) samples. Sampling year not reported.
n = 26, DF 73%, mean, median (range) = 0.32,
0.30 ( < ND-0.73) ng/g
(MDL = 0.05 ng/g)
Liu et al. (2014)
United States
(unspecified)
Consumer products commonly used indoors
were purchased between March 2007 and
September 2011 from local retailers and online
stores in the United States. A subset of samples
were analyzed for PFSAs and included
commercial carpet/fabric-care liquids,
household carpet/fabric-care liquids and
foams, treated apparel (i.e., one girl's uniform
pants and one boy's uniform pants), treated
home textile and upholstery (i.e., mattress
pads), treated floor waxes and stone-wood
sealants, membranes for apparel, and thread-
sealant tapes and pastes. The subset of
products originated from the United States,
England, Dominican Republic, Vietnam, and
China.
Commercial carpet/fabric-care liquids: n = 2,
DFa = 50%, range = BDL-194 ng/g
Household carpet/fabric-care liquids and foams:
n = 4, DFa = 50%, range = BDL-155 ng/g
Treated apparel: n = 2, DFa = 50%,
range = BDL-1.70 ng/g
Treated home textile and upholstery: n = 2,
DFa = 50%, range = BDL-12.1 ng/g
Treated floor waxes and stone-wood sealants:
n = 1, DF 0%
Membranes for apparel: n = 1, point = 7.10 ng/g
Thread-sealant tapes and pastes: n = 2,
DFa = 50%, range = BDL-60.3 ng/g
(DL not reported)
Notes: BDL = below detection limit; DF = detection frequency; DL = detection limit; MDL = method detection limit ND = not detected.
a The DF and/or mean was not reported in the study and was calculated in this synthesis. Means were calculated only when DF = 100%.
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Indoor Dust
In a Wisconsin Department of Health Services study, Knobeloch et al. (2012) examined levels of
16 perfluoroalkyl chemicals in vacuum cleaner dust from 39 Wisconsin homes across 16
counties in March and April 2008 (Table B-5). Samples from these homes built between 1890
and 2005 were collected during a pilot study to assess residential exposure to persistent
contaminants found in the Great Lakes Basin. PFHxS was found in all samples at a median
concentration of 16 ng/g. Mean levels of PFHxS in dust were significantly higher in homes built
between 1968 and 1995 (219 ng/g vs. 57 ng/g in homes constructed in other years). Based on the
results of this study, the authors suggested that perfluoroalkyl chemicals may be ubiquitous
contaminants in U.S. homes. In an EPA study of 112 indoor dust samples collected from vacuum
cleaner bags from homes and daycare centers in North Carolina and Ohio in 2000-2001 (EPA's
CTEPP study), samples were collected from 102 homes and 10 daycare centers in North
Carolina (49 homes, 5 daycare centers) and Ohio (53 homes, 5 daycare centers) (Strynar and
Lindstrom, 2008). Results were not reported separately for homes and daycares. Overall, PFHxS
was detected in 77.7% of all samples (n = 112) at mean and median concentrations of 874 and
45.5 ng/g, respectively. The authors concluded that the study measured perfluorinated
compounds in house dust at levels that may represent an important pathway for human exposure.
Additional peer-reviewed studies were identified that evaluated the occurrence of PFHxS and
other PFAS in dust of indoor environments, primarily in homes, as well as in schools, childcare
facilities, offices, and vehicles (Byrne et al., 2017; Fraser et al., 2013; Karaskova et al., 2016;
Kato et al., 2009; Knobeloch et al., 2012; Scher et al., 2019; Wu et al., 2014; Zheng et al., 2020)
(Table B-5). For those studies with results stratified for U.S. homes, PFHxS levels and detection
frequencies were lowest in a study of remote Alaska Native villages (27% detection, median
below 0.2 ng/g), while in other U.S. locations, PFHxS was detected in at least 40% of samples
(some studies reporting 100% detection) at widely varying mean and median levels across the
studies (from on the order of 10 ng/g to on the order of 200 ng/g) with one study reporting the
highest mean value (219 ng/g) from homes built between 1968 and 1995. The two studies also
reporting home measurements from other countries differed in how PFHxS levels in the U.S.
ranked relative to other countries, with one study ranking the U.S. highest and the other second
lowest. Few studies sampled childcare centers, vehicles, and offices, and none of the reviewed
studies reported measurements in other microenvironments (e.g., public libraries, universities).
Several studies reported results from dust samples collected only from homes (Byrne et al., 2017;
Scher et al., 2019; Wu et al., 2014), with one study sampling from locations near a PFAS
production facility. Scher et al. (2019) evaluated indoor dust in 19 homes in Minnesota within a
GCA in the vicinity of a former 3M PFAS production facility. Homes within the GCA had
previous or ongoing PFAS contamination in drinking water and were served by the Oakdale,
Minnesota public water system or a private well previously tested and shown to have detectable
levels of PFOA or PFOS. In the house dust samples, collected from July to September 2010, the
detection frequencies for PFHxS were 68% and 84% for entryways to the yard and interior living
spaces such as the family or living rooms, respectively (n = 19 each), with median concentrations
of 8.2 ng/g and 18 ng/g, respectively. PFAS concentrations in both sampling locations were
higher than corresponding soil concentrations, suggesting that interior sources were the main
contributors to PFAS in house dust.
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Byrne et al. (2017) assessed exposure to PFHxS and other PFAS among residents of two remote
Alaska Native villages on St. Lawrence Island. PFAS concentrations were measured in dust
collected from the surfaces of floors and furniture of 49 homes on St. Lawrence Island during
February-April of 2013 and 2014. Residents were asked not to sweep or dust for one week prior
to sampling. The authors described the overall PFAS levels in dust samples as "on the lower end
of those reported worldwide in other studies." PFHxS was found in 27% of all samples (n = 49)
with a median value below the LOD (0.1 ng/g-0.2 ng/g). Wu et al. (2014) measured
concentrations of five PFAS in residential dust in California in 2008-2009. Dust samples were
collected from the carpet or area rug in the main living area of the home. Homes of parents with
young children and homes with older adults were differentiated to characterize the relationship
between serum concentrations of PFAS and several other factors, including PFAS concentrations
in residential dust. PFHxS was detected in 51% of samples from households with young children
in Northern California (n = 82), with mean and median concentrations of 142 ng/g and 5.30 ng/g,
respectively. PFHxS was detected in 52% of samples from households of older adults in central
California (n = 42), with mean and median concentrations of 55 ng/g and 5.55 ng/g, respectively.
Apart from the information reported by Strynar and Lindstrom (2008), one other study included
childcare centers in the locations sampled, (Zheng et al., 2020). Zheng et al. (2020) collected
dust samples from seven childcare centers in Seattle, Washington (n = 14) and one childcare
facility in West Lafayette, Indiana (n = 6 across six rooms); the sampling year was not reported.
The included childcare facilities consisted of several building types, including multiple
classrooms, a former church, and a former home. Because centers were vacuumed and mopped
daily, dust samples were obtained from elevated surfaces (shelving, tops of bookcases/storage
cubbies) along with floor dust. PFHxS was detected in 95% of samples at mean and median
concentrations of 0.34 ng/g and 0.25 ng/g, respectively.
One study evaluated PFHxS levels in vehicles and offices, in addition to homes. Fraser et al.
(2013) collected dust samples between January and March 2009 from 3 microenvironments of 31
individuals in Boston, Massachusetts (offices (n = 31), homes (n = 30), and vehicles with
sufficient dust for analysis (n = 13)). Study participants worked in separate offices located across
seven buildings, which were categorized as Building A (n = 6), Building B (n = 17), or Other
(n = 8). Building A was a newly constructed (approximately one year prior to study initiation)
building with new carpeting and new upholstered furniture in each office. Building B was a
partially renovated (approximately one year prior to study initiation) building with new carpeting
throughout hallways and in about 10% of offices. The Other buildings had no known recent
renovation occurred. Study offices were not vacuumed during the sampling week and
participants were asked not to dust or vacuum their homes and vehicles for at least one week
prior to home sampling. Because PFHxS was detected in less than 50% of samples in all three
microenvironments, geometric means were not reported. The detection frequencies for PFHxS
were 23%, 40%, and 46% for offices, homes, and vehicles, respectively, with the range of
detected values reported as 5.24 ng/g-18.5 ng/g, 6.05 ng/g-430 ng/g, and 5.22 ng/g-108 ng/g,
respectively.
Two studies evaluated dust samples collected across multiple continents (Karaskova et al., 2016;
Kato et al., 2009). Karaskova et al. (2016) examined PFAS levels in house dust collected
between April and August 2013 from the living rooms and bedrooms of 14 homes in the United
States, 15 homes in Canada, and 12 homes in the Czech Republic (locations unspecified). PFHxS
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was detected in all U.S. samples (n = 20) at mean and median concentrations of 13.8 ng/g and
8.7 ng/g, respectively. The authors reported significant differences between countries for PFHxS
concentrations, with a trend of U.S. > Canada ~ Czech Republic and suggested that the
differences may be explained by differences in the market, import history, and usage of these
substances. (Note: As stated previously, while studies from non-U.S. countries inform an
understanding global exposure sources and trends, the RSC determination is based on the
available data for the United States). Overall, no significant differences in total PFAS
concentrations were found between the bedroom and living room in the same household
although significant relationships were found based on type of floors, number of residents, and
age of the house. A second multicontinental study (Kato et al., 2009) measured PFC
concentrations in 39 household dust samples collected in 2004 from homes in the United States
(Atlanta, GA) (n = 10), United Kingdom (n = 9), Germany (n = 10), and Australia (n = 10).
Across all 39 homes, PFHxS was detected in 79.5% of samples with a median concentration of
185.5 ng/g. Authors presented the median and maximum PFHxS concentrations by country in a
bar chart, which showed that PFHxS was detected in all countries. The median and maximum
PFHxS concentrations for the 10 United States (Atlanta, GA) house dust samples were
approximately 96.4 ng/g and 231.3 ng/g, respectively. The highest median was found in
Australia, followed by the United Kingdom, the United States, and Germany in decreasing order;
statistical significance on the comparison of median PFHxS concentrations by country was not
reported.
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Table B-5. Summary of PFHxS Indoor Dust Data
Study
Location
Site Details
Results
United States
Scheretal. (2019)
United States (Twin
Cities metropolitan
region, Minnesota)
Nineteen homes in three cities within a GCA near
former 3M PFAS production facility as well as from
three homes in the Twin Cities Metro outside the
GCA. Dust samples collected from an entryway to
the yard and from an interior living space (e.g.,
family room, living room) in each home in July-
September 2010. Homes within the GCA had
previous or ongoing PFAS contamination in drinking
water and were served by the Oakdale, Minnesota
public water system or a private well previously
tested and shown to have detectable levels of PFOA
or PFOS. Results were not reported for homes
outside the GCA.
Entryway: n = 19, DF 68%, median
(range) = 8.2 (< RL-94) ng/g
Living room: n = 19, DF 84%, median
(range) = 18 (< RL-790) ng/g
(RL = 5 ng/g)
Byrne et al. (2017)
United States (St.
Lawrence Island, Alaska)
Dust samples collected from the surfaces of floors
and furniture from 49 homes during February-April
of 2013 and 2014. Participants were asked not to
sweep or dust for one week prior to sampling.
n = 49; DF 27%, median (95th
percentile) = < LOD (3.13) ng/g
(MDL = 0.1-0.2 ng/g for all PFAS)
Wu et al. (2014)
United States (Central
Valley area, California)
Distributions of PFC dust concentrations were
determined for households with young children in
Northern California (n = 82) and households of older
adults in central California (n = 42). Dust samples
were collected in 2008-2009 from the carpet or area
rug in the main living area of the homes. Homes of
parents with young children and homes with older
adults were differentiated to characterize the
relationship between serum concentrations of PFCs
and PFC concentrations measured in residential dust.
Parents of young children: n = 82,
DF 51%, mean, median
(range) = 142, 5.30 (ND-7,490) ng/g
Older adults: n = 42, DF 52%, mean,
median (range) = 55, 5.55 (ND-
1,050) ng/g
(LOD = 0.10 ng/ml)
*Data below LOQ replaced by
LOD/V2
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Study
Location
Site Details
Results
United States
Dust samples were collected by the Wisconsin n = 39, DF 100%, median (range) = 16
Department of Health Services from 39 Wisconsin (2.1-1,000) ng/g
homes across 16 counties in March-April 2008. (rl = 1 ng/g)
Vacuum bags were collected or bagless vacuums
were emptied into sterilized glass jars. Homes were
built between 1890 and 2005.
Seven childcare centers in Seattle (14 samples) and n = 20; DF 95%, mean, median
one center in Lafayette (6 samples); sampling year (range) = 0.34, 0.25 (< ND-0.89) ng/g
not reported. Since all centers were vacuumed and (MDL = 0 05 ng/g)
mopped daily, dust samples from elevated surfaces
(shelving, tops of bookcases/storage cubbies) were
collected along with floor dust in the same sample.
Strynar and United States (North Dust samples from vacuum cleaner bags were
Lindstrom (2008) Carolina; Ohio) obtained in 2000-2001 during EPA's Children's
Total Exposure to Persistent Pesticides and Other
Persistent Organic Pollutants (CTEPP) study from
North Carolina (49 homes, 5 daycare centers) and
Ohio (53 homes, 5 daycare centers). Vacuum cleaner
bags were only collected if available at each site.
Knobeloch et al. United States (Great
(2012) Lakes Basin, Wisconsin)
Zheng et al. (2020)
United States (Seattle,
Washington; West
Lafayette, Indiana)
n = 112, DF 77.7%, mean, median
(maximum) = 874, 45.5 (35,700) ng/g
(LOQ = 12.9 ng/g)
*Values below the LOQ assigned a
value of LOQ/V2
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Study
Location
Site Details
Results
United States
Fraser et al. (2013)
United States (Boston,
Massachusetts)
Dust samples were collected between January-
March 2009 from offices (n = 31), homes (n = 30),
and vehicles (n = 13) of 31 individuals. Study
participants worked in separate offices located across
seven buildings, which were categorized into
Building A, Building B, and Other. Six samples were
collected from Building A, a newly constructed
(approximately one year prior to study initiation)
building with new carpeting and new upholstered
furniture in each office. Seventeen samples were
collected from Building B, a partially renovated
(approximately one year prior to study initiation)
building with new carpeting throughout hallways and
in about 10% of offices. Eight samples were
collected from the other five remaining buildings
where no known recent renovation occurred. Study
offices were not vacuumed during the sampling week
and homes and vehicles were not vacuumed for at
least one week prior to sampling. Entire accessible
floor surface areas and tops of immovable furniture
were vacuumed in offices and the main living area of
homes. Entire surface areas of the front and back
seats of vehicles were vacuumed.
Number of home dust samples was reduced to 30
because 1 participant lived in a boarding house with
no main living area. Sufficient mass of dust for
analysis was available from only 13 vehicles.
Homes: n = 30, DF 40%,
range = 6.05-430 ng/g
Offices: n = 31, DF 23%,
range = 5.24-18.5 ng/g
Vehicles: n = 13, DF 46%,
range = 5.22-108 ng/g
(LOQ = 5 ng/g)
*Range of detected values reported
Notes: DF = detection frequency; GCA = groundwater contamination area; LOD = limit of detection; LOQ = limit of quantitation; MDL = method detection limit; MQL = method
quantification limit; ND = not detected; RL = reporting limit.
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Air
Perfluoroalkyl chemicals have been released to air from wastewater treatment plants, waste
incinerators, and landfills (EPA, 2016), though there is limited information on the detection
levels or frequencies of PFHxS in either indoor or ambient air. NCBI (2022) notes that PFHxS
has been detected in particulates and in the vapor phase in air and it can be transported long
distances via the atmosphere; it has been detected at low concentrations in areas as remote as the
Arctic and ocean waters. For example, PFHxS was detected in particle-phase air samples
collected in 2007 and 2008 from the Atlantic Ocean, Southern Ocean, and Baltic Sea (Dreyer et
al., 2009; NCBI, 2022). PFHxS is not expected to be broken down directly by photolysis (NCBI,
2022). PFHxS in the particle-phase can be removed from the atmosphere through wet and dry
deposition (NCBI, 2022). PFHxS in the vapor phase can undergo hydroxylation in the
atmosphere, with a (predicted average) atmospheric hydroxylation rate of 2.16 x 10"
15 cmVmolecule - second to a (derived) rate of 1.4 x 10"13 cmVmolecule - second at 25 °C (with
corresponding estimated half-life of 115 days for this reaction in air) (NCBI, 2022; EPA, 2022a).
With a vapor pressure of 0.0046 mm Hg at 25 °C (estimated), 8.10 x 10"9 mm Hg (measured
average), and 8.19 x 10"9 mm Hg (predicted average), volatilization is not expected to be an
important fate process for this chemical (NCBI, 2022, EPA, 2022a). EPA's Toxics Release
Inventory reported release data for PFHxS in 2020, with total on-site disposal, off-site disposal,
and other releases concentrations of 1 pound at an individual facility and 122 pounds at a second
facility (EPA, 2022b). PFHxS is not listed as a hazardous air pollutant (EPA, 2022c).
indoor Air
No studies from the U.S. reporting levels of PFHxS in indoor air were identified from the
primary or gray literature.
Ambient Air
Kim and Kannan (2007) analyzed particle phase (n = 8) and gas phase (n = 8) concentrations of
perfluorinated acids in ambient air samples collected in and around Albany, New York in May
and July 2006 to examine the relative importance of certain media pathways to the contamination
of urban lakes. PFHxS was detected in all gas phase samples with mean, and median
concentrations of 0.31 pg/m3 and 0.34 pg/m3, respectively, but was not detected in the particle
phase (LOQ = 0.12 pg/m3).
Soil
The production of PFHxS and its use as a raw material or precursor for manufacturing PFAS
based products, as well as its previous use in firefighting foam and carpet treatment solutions,
could result in its release to soils through various waste streams (NCBI, 2022). When released to
soil, PFHxS is expected to have very high mobility (NCBI, 2022). The maximum concentration
of PFHxS in soil samples collected from a PFAS production facility in Minnesota was
3,470 ng/g, with PFHxS detected in 90% of the samples collected (3M, 2007; ATSDR, 2021;
NCBI, 2022). The maximum concentration of PFHxS in soil samples collected at a fire training
area at a PFAS production facility in Minnesota was 62.2 ng/g, with PFHxS detected in 100% of
samples (3M, 2007; ATSDR, 2021). PFHxS was also detected in soil samples collected from a
former sludge incorporation area at a PFAS production facility in Decatur, Alabama, with
average levels ranging from 3.56 ng/g to 270 ng/g, with PFHxS detected in 86% of the samples
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collected (3M, 2012; ATSDR, 2021). PFHxS has been found to accumulate in the roots of maize
plants grown in soil containing PFHxS and other PFAS (Krippner et al., 2015; ATSDR, 2021).
Seven additional peer-reviewed studies were identified that evaluated the occurrence of PFHxS
and other PFAS in soil (Anderson et al., 2016; Blaine et al., 2013; Eberle et al., 2017; Nickerson
et al., 2020; Scher et al., 2018, 2019; Venkatesan and Halden, 2014) (Table B-6). Among the
studies conducted in the United States, three analyzed soils potentially impacted by past AFFF
use. At all such sites, PFHxS was detected in at least 50% of samples, typically at levels less than
100 ng/g, but some values were higher and, in some cases, the subsurface PFHxS levels were
higher than in the surface soil. Other than control soils in two greenhouse and field studies and
one reference site, the U.S. studies did not evaluate soils without amendments or without a
nearby current or historical PFAS source.
Two studies by Scher et al. (2018; 2019) evaluated soils collected in 2010 from the garden
planting area of 20 homes in Minnesota within a GCA impacted by the former 3M PFAS
production facility. Homes within the GCA had previous or ongoing PFAS contamination in
drinking water and were served by the Oakdale, Minnesota public water system or a private well
previously tested and shown to have detectable levels of PFOA or PFOS. Both studies reported
similar median PFHxS levels of 0.08 ng/g and 0.057 ng/g (n = 20-34) from the 2019 and 2018
publications, respectively. Scher et al. (2018) also reported a median PFHxS concentration of
0.078 ng/g from six samples collected outside the GCA.
Three studies analyzed soils potentially impacted by AFFF use (Anderson et al., 2016; Eberle et
al., 2017; Nickerson et al., 2020). Anderson et al. (2016) examined surface and subsurface soil
from 40 sites across 10 active Air Force installations throughout the continental United States
and Alaska between March and September 2014. Installations were included if there was known
historic AFFF release in the period 1970-1990. It is assumed that the measured PFAS profiles at
these sites reflect the net effect of several decades of all applicable environmental processes. The
selected sites were not related to former fire training areas and were characterized according to
volume of AFFF release—low, medium, and high. Across all sites, the PFHxS detection
frequency was 76.92% in 100 surface soil samples (median concentration of detects was
5.7 ng/g) and 59.62%> in 112 subsurface soil samples (median concentration of detects was
4.4 ng/g). PFHxS was detected more frequently at high-volume release sites (82.5%> in 32
surface soil samples with mean concentration of 19.7 ng/g; 87.5%> in 31 subsurface soil samples
with mean concentration of 9.3 ng/g) than at low-volume sites (75.0%> in 12 surface soil samples
with mean concentration of 13.9 ng/g; 58.8%> in 17 subsurface soil samples with mean
concentration of 57.9 ng/g) and medium-volume sites (59.2%> in 56 surface soil samples with
mean concentration of 39.4 ng/g; 71.4% in 64 subsurface soil samples with mean concentration
of 55.4 ng/g). Nickerson et al. (2020) developed a method to quantify anionic, cationic, and
zwitterionic PFAS from AFFF-impacted soils. The method was applied to two soil cores
collected from two different AFFF-impacted former fire training areas; the sampling year and
geographic location were not provided. Eleven soil samples, corresponding to 11 depths ranging
from 0.46 m to 15.1 m, were evaluated from Core E, and 12 soil samples, at depths ranging from
0.30 m to 14.2 m, were evaluated from Core F. PFHxS was detected at all depths analyzed for
both cores, with concentrations ranging from 1.17 ng/g dw to 160.6 ng/g dw for Core E and
0.66 ng/g dw to 296.4 ng/g dw for Core F. Eberle et al. (2017) investigated the effects of an in
situ chemical oxidation treatment for remediation of chlorinated volatile organic compounds and
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PFAAs co-contaminants. Soil samples were collected in 2012-2013 before and after a pilot scale
field test at a former fire training site at Joint Base Langley-Eustis, Virginia. Monthly fire
training activities were conducted at the site from 1968 to 1980 and irregular fire training
activities continued until 1990. Impacted soil was excavated in 1982 but details were not
provided. PFHxS was detected in 4 of 5 pre-treatment samples and in all 14 post-treatment
samples. In the available three paired pre- and post-treatment soil samples, two pairings showed
a decrease in PFHxS concentration after treatment, from 6.7 ng/g to 1.40 ng/g in one pairing and
from 12 ng/g to 1.44 ng/g in the other pairing. In the third pairing, PFHxS was not detected pre-
treatment and was found at 0.31 ng/g post-treatment.
Of the remaining two studies conducted in the United States, Venkatesan and Halden (2014)
conducted outdoor mesocosm studies to examine the fate of PFAS in biosolids-amended soil
collected during 2005-2008. Biosolids were obtained from a WWTP in Baltimore that primarily
treated wastewater from domestic sources with only minor contribution (1.9%) from industry.
PFHxS was not detected in the control (nonamended) soil and not consistently detected in the
biosolids-amended soil (MDL = 0.03-0.14 ng/g dw). In a field and greenhouse study, Blaine et
al. (2013) studied the uptake of PFAS into edible crops grown in control and biosolids-amended
soil. In the field study, urban biosolids were obtained from a WWTP receiving both domestic
and industrial waste while rural solids were obtained from a WWTP receiving domestic waste
only. Mean PFHxS concentrations were below the LOQ (0.01 ng/g) in the urban control and
biosolids-amended soils and in the rural control soil. In the rural biosolids-amended soil, the
mean PFHxS concentration was 0.016 ng/g. In the greenhouse study, three soils (nonamended
control, industrially impacted, and municipal) were investigated. Industrially impacted soils
contained composted biosolids from a small municipal WWTP that was impacted by PFAA
manufacturing while municipal soils were obtained from a reclamation site in Illinois where
municipal biosolids were applied for 20 years. PFHxS was detected in all three soils at an
average concentration of 0.44 ng/g, 1.38 ng/g, and 5.11 ng/g in control, industrially impacted,
and municipal soil, respectively. Authors noted that the trace levels of PFAS detected in the
control soil may be due to minor cross-contamination from plowing, planting, or atmospheric
deposition from the surrounding area where biosolids have been applied.
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Table B-6. Summary of PFHxS Data in Soil
Study
Location
Site Details
Results
United States
Scher etal.
United States
Area near former 3M PFAS production facility.
n = 20, DF 85%, median (90th percentile) = 0.08
(2019)
(Twin Cities
Soil composite samples (200-500 g) collected in
(0.16) ng/g
metropolitan
2010 from the garden planting area of 20 homes
(RL = 0.0008-0.033 ng/g for all PFAS)
region,
in 3 cities within a GCA as well as from 3 homes
Minnesota)
in the Twin Cities Metro outside the GCA.
Homes within the GCA had previous or ongoing
PFAS contamination in drinking water and were
served by the Oakdale, Minnesota public water
system or a private well previously tested and
shown to have detectable levels of PFOA or
PFOS. Results were not reported for homes
outside the GCA.
Scher et al. United States
(2018) (Twin Cities
metropolitan
region,
Minnesota)
Area near former 3M PFAS production facility.
Soil samples collected in 2010 at a sample depth
of 0-6 inches from the garden planting area of 20
homes in 3 cities within a GCA as well as from 3
homes in the Twin Cities Metro outside the
GCA. Homes within the GCA had previous or
ongoing PFAS contamination in drinking water
and were served by the Oakdale, Minnesota
public water system or was formerly or currently
using a private well previously tested and shown
to have detectable levels of PFOA or PFOS. At
14 homes, soil samples were collected on two
separate days.
Within GCA: n = 34, DF 71%, median (range) = 0.057
(ND-0.24) ng/g
Outside GCA: n = 6, DF 100%, median (range) = 0.078
(0.028-0.11) ng/g
(MDL = 0.008-0.033 ng/g for all PFAS)
*Values below the method reporting limit but above the
lowest calibration limit (estimated values) were
included in all analyses as quantitative results
*Values below the lowest calibration limit were
replaced with !/2 of the limit
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Study
Location
Site Details
Results
United States
Anderson et al.
(2016)
United States Forty AFFF-impacted sites from ten active U.S.
(national) Air Force installations with historic AFFF
release between 1970 and 1990 that were not
related to former fire training areas. It is assumed
that the measured PFAS profiles at these sites
reflect the net effect of several decades of all
applicable environmental processes. AFFF-
impacted sites included emergency response
locations, hangers and buildings, and testing and
maintenance related to regular maintenance and
equipment performance testing of emergency
vehicles and performance testing of AFFF
solution. Previous remedial activities for co-
occurring contaminants were not specifically
controlled for in the site selection process; active
remedies had not been applied at any of the sites
selected. Approximately ten samples were
collected between March and September 2014 at
each site for surface and subsurface soil; sites
were grouped according to volume of AFFF
release—low-volume typically had a single
AFFF release, medium-volume had one to five
releases, and high-volume had multiple releases.
Surface soil:
Overall: n = 100, DF 76.92%, median
(maximum) = 5.7 (1,300) ng/g
Breakdown by site:
Emergency Response (low-volume release):
n = 12, DF 75.0%, mean (range) = 13.9 (0.38-
64) ng/g
Hangars and Buildings (medium-volume release):
n = 56, DF 59.2%, mean (range) = 39.4 (0.34-
1,300) ng/g
Testing and Maintenance (high-volume release):
n = 32, DF 82.5%, mean (range) = 19.7 (0.29-
180) ng/g
(RL = 0.29 ng/g)
Subsurface soil:
Overall: n = 112, DF 59.62%, median
(maximum) = 4.4 (520) ng/g
Breakdown by site:
Emergency Response (low-volume release):
n = 17, DF 58.8%, mean (range) = 57.9 (0.37-
520) ng/g
Hangars and Buildings (medium-volume release):
n = 64, DF 71.4%, mean (range) = 55.4 (0.35-
1,300) ng/g
Testing and Maintenance (high-volume release):
n = 31, DF 87.5%, mean (range) = 9.3 (1.1-40) ng/g
(RL = 0.31 ng/g)
* Median calculated using quantified detections
*Non-detects were substituted with !/2 the reporting limit
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Study
Location
Site Details
Results
United States
Nickerson et United States Soil cores E and F from two different AFFF- Core E:
al. (2020) (unspecified) impacted fire training areas; sampling year and 0.46 m = 1.44 ng/g dw
geographic location not provided. Soil core E 2.9 m = 2.12 ng/g dw
contained 11- 0.3 m increment samples from 3.66 m = 4.17 ng/g dw
0.3-15.2 m below ground surface and was 3.96 m = 15.21 ng/g dw
collected in an area where the surficial soils were 4.27 m = 28.68 ng/g dw
likely disturbed due to regrading and other soil 4.57 m = 4.13 ng/g dw
redistribution activities. Soil core F contained 4.88 m = 5.73 ng/g dw
12- 0.61 m increment samples from 0-14.2 m 7.01 m = 13.86 ng/g dw
below ground surface and was collected in an 8.38 m = 160.6 ng/g dw
area where the surficial soils were highly 10.5 m = 139.0 ng/g dw
permeable only within the upper 0.5 to 1 m, and 15.1m=1.17 ng/g dw
the underlying impermeable clay layer exhibited Core p
a relatively high cation exchange capacity and 030m=1107 ng/g dw
organic carbon content. The water table was 1 22 m = 296 4 ng/g dw
relatively shallow (depth < 3 m) at both sites. 1 83 m = 276 2 ng/g dw
2.44 m = 106.2 ng/g dw
3.05 m = 42.69 ng/g dw
4.11 m = 28.78 ng/g dw
7.62 m = 14.19 ng/g dw
8.84 m = 6.26 ng/g dw
9.45 m = 3.25 ng/g dw
10.5 m = 0.66 ng/g dw
11.9 m = 3.06 ng/g dw
14.2 m = 7.96 ng/g dw
(LOD/LOQ not reported)
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Study
Location
Site Details
Results
United States
Eberle et al.
(2017)
United States
(Joint Base
Langley-Eustis,
Virginia)
Pilot testing area in former fire training area
(Training Site 15) at Joint Base Langley-Eustis
where monthly fire training activities were
conducted from 1968 to 1980 in a zigzag pattern
burn pit. Facility was abandoned in 1980 but
irregular fire training activities using an above-
ground germed burn pit continued until 1990.
Impacted soil was removed in 1982 but
additional details of the excavation are not well
known. Soil samples collected for pre- (April
and September 2012) and post- (December 2013)
in situ chemical oxidation treatment using a
peroxone activated persulfate (OxyZone)
technology. Treatment was conducted in Test
Cell 1 over 113 days (April-August 2013). Soil
samples were collected adjacent to wells; wells
outside Test Cell 1 were used as sentry wells.
Well IDs for pre- and post-sampling were not
provided but the following three pairings were
assumed based on Table 2 in the paper: U-20
with SB-106; U-16 with SB-112; and 1-1 with
SB-109.
Pre-treatment:
1-1 (1.2-4.3 m) = 12 ng/g
1-2(1.2-4.3 m) = 83 ng/g
U-12 (2.1 m) = 1.2 ng/g
U-16 (3.0 m) = 6.7 ng/g
U-20 (1.8 m)=ND
(LOQ = 0.68-0.72 ng/g)
Post-treatment:
SB-101 (4.3 m) = 8.08 ng/g
SB-105 (1.8 m) = 0.83 ng/g
SB-106/U-20 (1.8 m) = 0.31 ng/g
SB-106 (4.3 m) = 5.08 ng/g
SB-107 (1.8 m) = 2.11 ng/g
SB-107 (4.3 m) = 3.99 ng/g
SB-108 (1.8 m) = 1.48 ng/g
SB-108 (4.3 m) = 4.83 ng/g
SB-109/1-1 (3 m) = 1.44 ng/g
SB-Ill (4.3 m) = 11.85 ng/g
SB-112 (1.8 m) = 2.57 ng/g
SB-112/U-16 (3 m) = 1.4 ng/g
SB-114 (1.8 m) = 3.63 ng/g
SB-114 (4.3 m) = 16.17 ng/g
(LOQ = 0.12 ng/g)
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Study
Location
Site Details
Results
United States
Venkatesan
and Halden
(2014)
United States
(Baltimore,
Maryland)
Archived agricultural soil (nonamended)
collected during 2005-2008 at a depth of 0-
20 cm from the United States Department of
Agriculture-Agricultural Research Service
Beltsville Agricultural Research Center; number
of sampling sites and number of samples not
provided.
Nonamended: n = NR, DF 0%
Amended: n = NR, DFa 0%
(MDL = 0.03-0.14 ng/g dw for all PFAS)
Biosolids-amended soil obtained by mixing
biosolids and soil at a volumetric ratio of 1:2.
Biosolids were from Back River WWTP in
Baltimore, a full-scale activated sludge treatment
plant. Raw wastewater was primarily from
domestic sources with only minor contribution
(1.9%) from industry.
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Study
Location
Site Details
Results
United States
Blaine et al. United States Urban and rural full-scale field study with
(2013) (Midwest) control (nonamended) and biosolids-amended
plots. Three agricultural fields were amended
(0.5x, lx, or 2x) with municipal biosolids.
Urban biosolids (1 x and 2x) were from a WWTP
receiving both domestic and industrial waste.
Rural biosolids (0.5 x) were from a WWTP
receiving domestic waste only. Control plots
were proximal to the rural and urban amended
corn grain and corn stover field sites; sampling
year not provided.
Greenhouse study with control (nonamended)
and biosolids-amended soil. Nonamended soil
obtained from a field that received commercial
fertilizers and had a similar cropping system as
the nearby municipal soil site. Municipal soil
was obtained from a reclamation site in Illinois
where municipal biosolids were applied at
reclamation rates for 20 years, reaching the
cumulative biosolids application rate of
1,654 Mg/ha. Industrially impacted soil was
created by mixing composted biosolids from a
small municipal (but impacted by PFAA
manufacturing) WWTP with control soil on a
10% mass basis. Sampling year not provided.
Field study:
Urban non-amended: n = 3-7, DF NR,
mean <0.01 ng/g
Urban lx; n = 3-7, DF NR, mean < 0.01 ng/g
Urban 2x: n = 3-7, DF NR, mean < 0.01 ng/g
Rural non-amended: n = 3-7, DF NR,
mean < 0.01 ng/g
Rural 0.5 x; n = 3-1, DF NR, mean = 0.16 ng/g
(LOQ = 0.01 ng/g)
Greenhouse study:
Nonamended: n = 3-5, DF NR, mean = 0.44 ng/g
Industrially impacted: n = 3-5, DF NR,
mean =1.38 ng/g
Municipal: n = 3-5, DF NR, mean = 5.11 ng/g
(LOQ not reported)
Notes: AFFF = aqueous film-forming foam; DF = detection frequency; dw = dry weight; GCA = groundwater contamination area; LOD = limit of detection; LOQ = limit of
quantitation; MDL = method detection limit; 0.5 x, lx, or 2x = !/2,1, or 2 times the agronomic rate of biosolids application to meet nitrogen requirements of the crop; ND = not
detected; NR = not reported; RL = reporting limit; WWTP = wastewater treatment plant.
a Venkatesan and Halden (2014) reported that PFHxS was not consistently detected in biosolids-amended mesocosms.
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Sediment
When released into water, PFHxS is not expected to adsorb to suspended solids and sediments
(NCBI, 2022). However, PFHxS was detected in 28% of sediment samples collected along the
Mississippi River shoreline in the vicinity of a PFAS production facility in Minnesota, at a
maximum concentration of 11.5 ng/g (3M, 2007; ATSDR, 2021; NCBI, 2022). PFHxS was also
detected in 96% of sediment samples collected from two coves of the Mississippi River (East
and West coves) located at the eastern and western ends of the PFAS production facility, at a
maximum concentration of 126 ng/g (3M, 2007; ATSDR, 2021). PFHxS was not detected in any
Mississippi River transect sediment samples (collected at points crossing the river near the PFAS
facility) (3M, 2007; ATSDR, 2021). PFHxS has been detected in sediment core samples
collected from three Canadian Arctic lakes in 2003 and 2005 at concentrations ranging from
approximately 1 ng/g to 10 ng/g (NCBI, 2022; Stock et al., 2007).
Biomonitoring in the U.S. Population
As indicated by CDC's NHANES results, PFHxS was detected in the blood of > 97% of
NHANES participants for most of the years in which it was evaluated (CDC, 2019, 2021a, b,
2022). Whole-weight serum levels of PFHxS in the 50th percentile of the U.S. population for all
years evaluated since 1999 were 2.10 |ig/L in 1999-2000 (detected in 99.7% of samples),
1.90 |ig/L in 2003-2004 (detected in 97.7% of samples), 1.80 |ig/L in 2005-2006 (detected in
95.9% of samples), 2.00 |ig/L in 2007-2008 (detected in 99.2% of samples), 1.70 |ig/L in 2009-
2010 (detected in 99.4% of samples), 1.27 |ig/L in 2011-2012 (detected in 98.4% of samples),
1.40 |ig/L in 2013-2014 (detected in 98.8% of samples), 1.20 |ig/L in 2015-2016 (detected in
98.4%) of samples), and 1.10 |ig/L in 2017-2018 (detected in 99% of samples) (CDC, 2019,
2021 a,b, 2022).
Recommended RSC
In summary, based on the physical properties, detected levels, and available exposure
information for PFHxS, multiple non-drinking water sources (fish and shellfish, non-fish food,
some consumer products, indoor dust, and soil) are potentially significant exposure sources.
Following the Exposure Decision Tree in EPA's 2000 Methodology (EPA, 2000), significant
potential sources other than drinking water ingestion were identified (Box 8A in the Decision
Tree); however, information is not available to quantitatively characterize exposure from these
different sources (Box 8B in the Decision Tree). Therefore, EPA recommends an RSC of 20%
(0.20) for PFHxS.
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