4>EPA
United States Office of Water EPA 822-R-16-005
Environmental Protection Mail Code 4304T May 2016
Agency
Drinking Water Health
Advisory for
Perfluorooctanoic Acid
(PFOA)
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) - May 2016
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Drinking Water Health Advisory
for Perfluorooctanoic Acid (PFOA)
Prepared by:
U.S. Environmental Protection Agency
Office of Water (43 04T)
Health and Ecological Criteria Division
Washington, DC 20460
EPA Document Number: 822-R-16-005
May 2016
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) - May 2016
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ACKNOWLEDGMENTS
This document was prepared by the Health and Ecological Criteria Division, Office of
Science and Technology, Office of Water of the U.S. Environmental Protection Agency (EPA).
The Agency gratefully acknowledges the valuable contributions of EPA scientists Glinda
Cooper, Ph.D.; Barbara Glenn, Ph.D.; Erin Hines, Ph.D.; Michael Wright, Sc.D.; John
Wambaugh, Ph.D.; Thomas Speth, Ph.D.; and Daniel Hautman.
This Health Advisory was provided for review by and comments were received from staff in
the following EPA program Offices:
Office of Chemical Safety and Pollution Prevention
Office of Children's Health Protection
Office of General Counsel
Office of Land and Emergency Response
Office of Policy
Office of Research and Development
Office of Water
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) - May 2016
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CONTENTS
ACKNOWLEDGMENTS 3
ABBREVIATIONS AND ACRONYMS 7
EXECUTIVE SUMMARY 9
1.0 INTRODUCTION AND BACKGROUND 11
1.1 Safe Drinking Water Act 11
1.2 Current Advisories and Guidelines 12
1.3 UsesofPFOA 14
2.0 NATURE OF THE STRESSOR 15
2.1 Physical and Chemical Properties 15
2.2 Occurrence and Sources of Exposure 17
2.2.1 Surface Water and Ground Water 17
2.2.2 Drinking Water 18
2.2.3 Food 18
2.2.4 Ambient Air 21
2.2.5 Indoor Dust 22
2.2.6 Soils 22
2.2.7 Biosolids 23
2.2.8 Consumer Products 24
2.3 Environmental Fate 24
2.3.1 Mobility 24
2.3.2 Persistence 24
2.3.3 Bioaccumulation 25
2.4 Toxicokinetics 25
2.5 Human Biomonitoring Data 27
3.0 PROBLEM FORMULATION 28
3.1 Conceptual Model 28
3.1.1 Conceptual Model Diagram for Exposure via finished Drinking Water 29
3.1.2 Factors Considered in the Conceptual Model forPFOA 29
3.2 Analysis Plan 31
3.2.1 Health Advisory Guidelines 31
3.2.2 Establishing the Data Set 31
3.2.3 Approach for HA Calculation 32
3.2.4 Measures of Effect 33
3.2.5 Relative Source Contribution 34
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4.0 EFFECTS ASSESSMENT 35
4.1 Noncancer Health Effects 35
4.1.1 Animal Toxicity Studies 35
4.1.2 Human Epidemiology Studies 39
4.1.3 Noncancer Mode of Action 42
4.2 Cancer 44
4.2.1 Animal Cancer Bioassays 44
4.2.2 Human Epidemiology Studies 45
4.2.3 Cancer Mode of Action 46
4.2.4 Weight of Evidence Classification 47
5.0 DOSE-RESPONSE ASSESSMENT 48
5.1 Uncertainty Factors 50
5.2 RfD Determination 51
6.0 HEALTH ADVISORY VALUES 53
6.1 Relative Source Contribution 53
6.2 Lifetime Health Advisory 54
7.0 QUANTIFICATION OF CANCER RISK 56
8.0 EFFECTS CHARACTERIZATION 57
8.1 Uncertainty and Variability 57
8.2 Use of Human Epidemiology Data 58
8.3 Consideration of Immunotoxicity 58
8.4 Effects on Mammary Gland Development 60
8.5 Alternative Exposure Scenarios 61
8.6 Relative Source Contribution Considerations 61
8.7 Sensitive Populations: Gender Differences 63
8.8 Sensitive Populations: Developmental Effects 63
9.0 ANALYTICAL METHODS 64
10.0 TREATMENT TECHNOLOGIES 65
11.0 REFERENCES 69
12.0 APPENDIX A-QUANTITATIVE CANCER ASSESSMENT MODELING 99
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TABLES
Table 1-1. State Guideline Values for PFOA 13
Table 1-2. International Guideline Values for PFOA 13
Table 2-1. Chemical and Physical Properties of PFOA 16
Table 5-1. Human Equivalent Doses Derived from the Modeled Animal Average Serum
Values 49
Table 5-2. Candidate RfDs Derived from the HEDs from the Pharmacokinetic Model
Average Serum Values 52
FIGURES
Figure 2-1. Chemical Structures of PFOA and APFO 15
Figure 3-1. Conceptual Model for PFOA in Finished Drinking Water 28
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ABBREVIATIONS AND ACRONYMS
ALT
ALP
APFO
AST
AUC
BAF
BCF
BMDL
BMP
bw
CAR
CCK
CCL
COPD
CWA
DWEL
DWI
EPA
FXR
GFR
GOT
HA
HDL
HED
HESD
IgM
IRIS
JS.OC
JVow
LCT
LC/MS/MS
LDL
LOAEL
MOA
MRL
ng/L
NHANES
NOAEL
PAC
PACT
PBPK
PFAS
PFC
PFOA
PFOS
alanine aminotransferase
alkaline phosphatase
ammonium perfluorooctanoate
aspartate aminotransferase
area under the curve
bioaccumulation factor
bioconcentration factor
benchmark dose level
biomagnification factor
body weight
constitutive androstane receptor
cholecystokinin
Contaminant Candidate List
chronic obstructive airways disease
Clean Water Act
drinking water equivalent level
drinking water intake
U.S. Environmental Protection Agency
farnesoid X receptor
glomerular filtration rate
gamma-glutamyl transferase
Health Advisory
high-density lipoprotein
human equivalent dose
Health Effects Support Document
immunoglobulin M
Integrated Risk Information System
organic carbon-water partitioning coefficient
octanol-water partition coefficient
Leydig cell tumor
liquid chromatography/tandem mass spectrometry
low-density lipoprotein
lowest observed adverse effect level
mode of action
minimum reporting level
nanograms per liter
National Health and Nutrition Examination Survey
no observed adverse effect level
powdered activated carbon
pancreatic acinar cell tumors
physiologically based pharmacokinetic model
perfluoroalkyl substance
perfluoroinated compounds
perfluorooctanoic acid
perfluorooctanesulfonic acid
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PTFE
Pg/L
PND
POD
POE
POU
PPARa
PWS
PXR
REACH
RfD
RSC
SDWA
SNUR
SRBC
IMF
TNSSS
UCMR3
UF
UV
polytetrafluoroethylene
picograms per liter
post-natal day
point of departure
point-of-entry
point-of-use
peroxisome proliferator activated receptor alpha
public water system
pregnane X receptor
Registration, Evaluation, Authorization, and Restriction of Chemicals
reference dose
relative source contribution
Safe Drinking Water Act
Significant New Use Rule
sheep red blood cell
trophic magnification factor
Total National Sewage Sludge Survey
third Unregulated Contaminant Monitoring Rule
uncertainty factor
ultraviolet
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EXECUTIVE SUMMARY
Perfluorooctanoic acid (PFOA) is a synthetic, fully fluorinated organic acid; it used in a
variety of consumer products and in the production of fluoropolymers, and it is generated as a
degradation product of other perfluorinated compounds. Because of strong carbon-fluorine
bonds, PFOA is stable to metabolic and environmental degradation. PFOA is one of a large
group of perfluoroalkyl substances (PFASs) that are used to make products more resistant to
stains, grease, and water. These compounds have been widely found in consumer and industrial
products, as well as in food items. Major U.S. manufacturers voluntarily agreed to phase out
production of PFOA by the end of 2015. Exposure to PFOA in the United States remains
possible due to its legacy uses, existing and legacy uses on imported goods, degradation of
precursors, and extremely high persistence in the environment and the human body. PFOA was
detected in blood serum in 99% of the U.S. general population between 1999 and 2012; however,
the levels of PFOA in blood have been decreasing since U.S. companies began to phase out
production. Water resources contaminated by PFOA have been associated with releases from
manufacturing sites, industrial sites, fire/crash training areas, and industrial or municipal waste
sites where products are disposed of or applied.
The U.S. Environmental Protection Agency (EPA) is issuing a lifetime drinking water Health
Advisory (HA) for PFOA of 0.07 micrograms per liter (|ig/L) based on a reference dose (RfD)
derived from a developmental toxicity study in mice; the critical effects included reduced
ossification in proximal phalanges and accelerated puberty in male pups following exposure
during gestation and lactation. PFOA is known to be transmitted to the fetus in cord blood and to
the newborn in breast milk. This lifetime HA is based on the latest health effects information for
noncancer and cancer effects for PFOA as described in EPA's 2016 Health Effects Support
Document for Perfluorooctanoic Acid (PFOA)., which was revised following external peer
review. Because the developing fetus and newborn are particularly sensitive to PFOA-induced
toxicity, the RfD based on developmental effects also is protective of adverse effects in adults
(e.g., liver and kidney toxicity). The lifetime HA is therefore protective of the population at
large.
For PFOA, oral animal studies of short-term, subchronic, and chronic duration are available
in multiple species including monkeys, rats and mice. These studies report developmental effects
(survival, body weight changes, reduced ossification, delays in eye opening, altered puberty, and
retarded mammary gland development), liver toxicity (hypertrophy, necrosis, and effects on the
metabolism and deposition of dietary lipids), kidney toxicity (weight), immune effects, and
cancer (liver, testicular, and pancreatic). Overall, the toxicity studies available for PFOA
demonstrate that the developing fetus is particularly sensitive to PFOA-induced toxicity. Human
epidemiology data report associations between PFOA exposure and high cholesterol, increased
liver enzymes, decreased vaccination response, thyroid disorders, pregnancy-induced
hypertension and preeclampsia, and cancer (testicular and kidney).
To derive candidate RfDs, EPA used a peer-reviewed pharmacokinetic model to calculate the
average serum concentrations associated with candidate no observed adverse effect levels
(NOAELs) and lowest observed adverse effect levels (LOAELs) from six studies for multiple
effects. Consistent with EPA's guidance A Review of the Reference Dose and Reference
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) - May 2016
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Concentration Processes (USEPA 2002), EPA applied protective uncertainty factors to address
intraspecies variability, interspecies variability, and LOAEL to NOAEL extrapolation.
From a national perspective, the dominant source of human exposure to PFOA is expected to
be from the diet; indoor dust from carpets and other sources also is an important source of
exposure, especially for children. The HA was calculated using a relative source contribution
(RSC) of 20%, which allows for other PFOA exposure sources (e.g., dust, diet, air) to make up
80%oftheRfD.
EPA's risk assessment guidelines reflect that, as a general matter, a single exposure to a
developmental toxin at a critical time in development can produce an adverse effect (USEPA
1991). In addition, short-term exposure to PFASs can result in a body burden that persists for
years and can increase with additional exposures. Thus, EPA recommends that the lifetime HA
for PFOA of 0.07 ug/L apply to both short-term (i.e., weeks to months) scenarios during
pregnancy and lactation, as well as to lifetime-exposure scenarios.
Adverse effects observed following exposures to PFOA and PFOS are the same or similar
and include effects in humans on serum lipids, birth weight, and serum antibodies. Some of the
animal studies show common effects on the liver, neonate development, and responses to
immunological challenges. Both compounds were also associated with tumors in long-term
animal studies. The RfDs for both PFOA and PFOS are based on similar developmental effects
and are numerically identical; when these two chemicals co-occur at the same time and location
in a drinking water source, a conservative and health-protective approach that EPA recommends
would be to compare the sum of the concentrations ([PFOA] + [PFOS]) to the HA (0.07
Under EPA's Guidelines for Carcinogen Risk Assessment (USEPA 2005), there is Suggestive
Evidence of Carcinogenic Potential for PFOA. Epidemiology studies demonstrate an association
of serum PFOA with kidney and testicular tumors among highly exposed members of the general
population. Two chronic bioassays of PFOA support a positive finding for the ability of PFOA to
be tumorigenic in one or more organs of rats, including the liver, testes, and pancreas. EPA
estimated a cancer slope factor of 0.07 per milligram per kilogram-day (mg/kg-day)"1 based on
testicular tumors, and confirmed that the lifetime HA based on noncancer effects is protective of
the cancer endpoint.
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1.0 INTRODUCTION AND BACKGROUND
The U.S. Environmental Protection Agency (EPA) developed the nonregulatory Health
Advisory (HA) Program in 1978 to provide information for public health officials or other
interested groups on pollutants associated with short-term contamination incidents or spills that
can affect drinking water quality, but are not regulated under the Safe Drinking Water Act
(SDWA). At present, EPA lists HAs for more than 200 contaminants.l
HAs identify the concentration of a contaminant in drinking water at which adverse health
effects are not anticipated to occur over specific exposure durations (e.g., 1 day, 10 days, a
lifetime). HAs serve as informal technical guidance to assist federal, state, and local officials,
and managers of public or community water systems in protecting public health when emergency
spills or other contamination situations occur. An HA document provides information on the
environmental properties, health effects, analytical methodology, and treatment technologies for
removing drinking water contaminants.
Perfluorooctanoic acid (PFOA) is a manmade chemical in a large family of chemicals called
perfluoroalkyl substances (PFASs) (Buck et al. 2011). PFOA has been used in a variety of
consumer products and in the production of fluoropolymers, and is generated as a degradation
product of other perfluorinated compounds. PFOA is very persistent in the environment and the
human body; it has been detected in water, wildlife, and humans worldwide. This document,
EPA's 2016 Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA), presents a
guideline concentration for PFOA in drinking water at which adverse health effects are not
anticipated to occur over a human lifetime. This lifetime HA is based on the latest health effects
information for noncancer and cancer effects for PFOA as described in EPA's Health Effects
Support Document for Perfluorooctanoic Acid (PFOA) (USEPA 2016a). The HA value is not a
legally enforceable federal standard and is subject to change as new information becomes
available. The structure, principles, and approach of this document are consistent with EPA's
Framework for Human Health Risk Assessment to Inform Decision Making (USEPA 2014a).
1.1 Safe Drinking Water Act
SDWA, as amended in 1996, requires EPA to publish a list of unregulated contaminants
every 5 years that are not subject to any proposed or promulgated national primary drinking
water regulations, are known or anticipated to occur in public water systems (PWSs), and might
require regulation under SDWA. This list is known as the Contaminant Candidate List (CCL).
PFOA is included on the third CCL (USEPA 2009a) and on the draft fourth CCL (USEPA
2015a).
As part of its responsibilities under SDWA, EPA is required to implement a monitoring
program for unregulated contaminants. SDWA requires, among other things, that once every
5 years, EPA issue a list of no more than 30 unregulated contaminants to be monitored by PWSs.
In 2012, EPA included PFOA in its third Unregulated Contaminant Monitoring Rule (UCMR 3),
which required all large systems serving > 10,000 people, plus a statistically selected group of
800 small systems to monitor for a 1-year period between 2013 and 2015. The last of the
1 For more information see http://water.epa.gov/drink/standards/hascience.cfm.
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) - May 2016 11
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monitoring data are still being compiled, but results to-date indicate that PFOA has been
measured at or above the minimum reporting level (0.02 micrograms per liter [ug/L]) by
approximately 2% of PWSs nationwide. To-date, PFOA has been measured above the new
lifetime HA level of 0.07 ug/L by approximately 0.3% of PWSs. Approximately 1% of PWSs
have reported data for which combined PFOA and PFOS results are above 0.07 ug/L. For the
latest UCMR 3 results, please refer to https://www.epa.gov/dwucmr/occurrence-data-
unregulated-contaminant-monitoring-rule#3.
SDWA requires EPA to make regulatory determinations for at least five CCL contaminants
every 5 years. EPA must begin developing a national primary drinking water regulation when the
Agency makes a determination to regulate based on three criteria:
• The contaminant may have an adverse effect on the health of persons.
• The contaminant is known to occur or there is substantial likelihood the contaminant will
occur in public water systems with a frequency and at levels of public health concern.
• In the sole judgment of the Administrator, regulating the contaminant presents a
meaningful opportunity for health risk reductions.
To make these determinations, the Agency uses data to analyze occurrence of these
compounds in finished drinking water and data on health effects. If EPA determines the
contaminant does not meet any one of the three statutory criteria, the Agency's determination is
not to regulate. EPA continues to gather information to inform future regulatory determinations
for PFOA under the SDWA.
EPA developed a Health Effects Support Document for Perfluorooctanoic Acid (PFOA) and
one for another PFAS, perfluorooctane sulfonate (also known as perfluorooctanesulfonic acid or
PFOS), to assist federal, state, tribal and local officials, and managers of drinking water systems
in protecting public health when these chemicals are present in drinking water (USEPA 2016a,
2016b). The health effects support documents (HESDs) were peer-reviewed in 2014 and were
revised as recommended by the peer reviewers with consideration of public comments and
inclusion of additional studies published through December 2015. The revised HESD for PFOA
(USEPA 2016a) provides an RfD and cancer assessment that serve as the basis for this HA.
The SDWA provides the authority for EPA to publish nonregulatory HAs or take other
appropriate actions for contaminants not subject to any national primary drinking water
regulation. EPA is providing this HA for PFOA to assist federal, state, and local officials
evaluate risks from this contaminant in drinking water. The HA values consider variability in
human response across all life stages and population groups while making allowance for
contributions from other exposure media.
1.2 Current Advisories and Guidelines
Currently there are no federal regulations under the SDWA or national recommended
ambient water quality criteria under the Clean Water Act (CWA) for PFOA. In January 2009,
EPA developed a provisional HA for PFOA in drinking water of 0.4 micrograms per liter (ug/L).
The provisional HA was developed to reflect an amount of PFOA that could cause adverse
health effects in the short term (weeks to months). The provisional HA was intended as a
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) - May 2016 12
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guideline for PWSs while allowing time for EPA to develop a lifetime HA. Table 1-1 provides
drinking water guideline values that were developed by states.
Table 1-1. State Guideline Values for PFOA
State
Delaware Department of Resources and Environmental Control
Maine Department of Health and Human Services
Michigan Department of Environmental Quality
Minnesota Department of Health
New Jersey Department of Environmental Protection
North Carolina Division of Water Quality
Vermont Agency of Natural Resources
Guideline Value
(MS/L)
0.4
0.1
0.42
0.3
0.04
2
0.02
Source
DNREC (2016)
Maine DHHS (2014)
Michigan DEQ (2013)
MDH (2009)
NJDEP (2014)
NCDEQ(2013)
Vermont ANR (20 16)
In 2013 the European Chemicals Agency adopted an agreement that identified PFOA as a
"Substance of Very High Concern" because of its persistent, bioaccumulative, and toxic
characteristics and placed it onto the Candidate List for Registration, Evaluation, Authorization
and Restriction of Chemicals (REACH) (Vierke et al. 2012). Once on the Candidate List, PFOA
could be included in Annex XIV of the REACH regulation, which would effectively ban use in
manufacturing and in the market.
PFOA also is being considered for listing under The Stockholm Convention on Persistent
Organic Pollutants (Convention), a global treaty to protect human health and the environment
from persistent organic pollutants. In October 2015, the Persistent Organic Pollutants Review
Committee agreed that PFOA meets the screening criteria in Annex D of the Convention, the
first of several steps toward listing of chemicals. Listing in various Annexes of the Convention
obligates parties to abide by provisions set forth to prohibit, eliminate, or restrict production and
use, as well as the import and export of persistent organic pollutants, except as allowed for by
specific exemptions. Several international agencies have established guideline values for PFOA
(Table 1-2).
Table 1-2. International Guideline Values for PFOA
Country/Agency
German Ministry of
Health
United Kingdom (UK)
Drinking Water
Inspectorate
Danish Ministry of the
Environment
Guideline Value (jig/ L)
Health-based
0.3
5.0
0.3
Administrative
Composite precautionary guidance value for
PFOA+PFOSisO.l
Action levels:
Tier 1 : potential hazard
Tier 2: > 0.3
Tier 3: > 5.0
Tier 4: > 45
Composite drinking water criteria are based
on relative toxicity of PFOS, PFOA, and
PFOSA
Source
German Ministry of
Health (2006)
UK Drinking Water
Inspectorate (2009)
Danish Ministry of the
Environment (2015)
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Country/Agency
Swedish National Food
Agency
Guideline Value (jig/ L)
Health-based
-
Administrative
Also 0.09 for the mixture of: PFOS, PFOA,
PFHxS; PFBS; PFHpA, PFHsA, PFPeA
(total PFASs)
0.9: Pregnant women, women trying to get
pregnant, and infants should not consume if
total PFASs exceeds
Source
Livsmedelsverket
(2014), cited in Danish
Ministry of the
Environment (2015)
Notes'.
PFOA = perfluorooctanoic acid; PFOS = perfluorooctane sulfonate; PFBS = perfluorobutane sulfonate; PFHpA =
perfluoroheptanoic acid; PFHsA = perfluorohexanoic acid; PFHxS = perfluorohexane sulfonic acid; PFOSA =
perfluorosulfonamide; PFPeA = perfluoropentanoic acid
1.3 Uses of PFOA
Perfluorinated substances, such as PFOA and its derivatives, are water- and lipid-resistant
because of their chemical properties. Therefore, they are commonly used as surface-active agents
that alter the surface tension of a mixture. Historically, PFOA was used in the United States in
carpets, leathers, textiles, upholstering, paper packaging, and coating additives as a
waterproofing or stain-resistant agent. Fire resistance of aviation fluid is increased by adding
PFOA, PFOS, and other PFASs to the mixture.
In 2006, EPA initiated the 2010/2015 PFOA Stewardship Program in which eight major
companies committed to reduce facility emissions and product contents of PFOA and related
chemicals on a global basis by 95% no later than 2010, and to work toward eliminating
emissions and product content of these chemicals by 2015 (USEPA 2006). Although the
2010/2015 PFOA Stewardship Program has worked toward eliminating emissions and product
content, there are still some ongoing uses that EPA is evaluating. Shorter-chain perfluoroalkyl-
based products have been developed to replace these chemicals.
To complement the Stewardship Program, EPA developed Significant New Use Rules2
(SNURs) to allow EPA to review any significant new uses of PFOA and many PFOA-related
chemicals before they are commercialized in the United States. On October 22, 2013, EPA
issued a final SNUR (published in the Federal Register [FR]; 78 FR 62443) requiring companies
to provide notice of any new manufacturing or processing of long-chain perfluoroalkyl
carboxylates for use in or on carpets (i.e., to impart soil, water, and stain resistance). Companies
must now provide EPA with notice of their intent to manufacture (including import) any of these
chemicals if they are used in carpets or to treat carpets. They must also notify EPA for these
chemical substances if they intend to import carpets already containing these chemical
substances. EPA subsequently proposed another SNUR on January 21, 2015, for PFOA and also
for PFOA-related chemicals that have not yet been commercialized (80 FR 2885).
2 For more information on EPA's SNURs visit http://www.epa.gov/assessing-and-managing-chemicals-under-
tsca/long-chain-perfluorinated-chemicals-pfcs.
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Given the limited ongoing uses of PFOA-related chemicals, releases to surface water and
ground water from PFOA are expected to decline. Exposure to PFOA in the United States
remains possible, however, because of its legacy uses, existing and legacy uses on imported
goods, degradation of precursors, and extremely high persistence in the environment and in the
human body.
2.0 NATURE OF THE STRESSOR
2.1 Physical and Chemical Properties
PFOA and its salts are fluorinated organic compounds and are part of the group of PFASs.
PFOA is a completely fluorinated organic synthetic acid that was used in the United States
primarily as an aqueous dispersion agent in the manufacture of fluoropolymers and in a variety
of water-, oil-, and stain-repellant products. Ammonium perfluorooctanoate (APFO) is the
ammonium salt of PFOA (Figure 2-1) which was a processing aid in the manufacture of certain
fluoropolymers, especially as an emulsifier during the polymerization of tetrafluoroethylene to
make polytetrafluoroethylene (e.g., Teflon™). Most of these primary uses have been voluntarily
phased out in the United States as of 2015 (see section 1.3 above); however, limited U.S. uses
and imports continue. Some sources of PFOA in the environment result from the atmospheric
degradation or transformation and/or surface deposition of precursors, including related
fluorinated chemicals (perfluorotelomer alcohols) (Wallington et al. 2006).
PFOA APFO
FFFFFFFF FFFFFFFF
Source: SIAR 2008
Figure 2-1. Chemical Structures of PFOA and APFO
The structure of PFOA varies with the manufacturing process. PFOA can be either a linear or
branched eight-carbon carboxylic acid with a partial negative charge on each fluorine and an
acidic carboxylate functional group. Low concentrations of other perfluorocarboxylate chain
lengths can also be present. It will tend to form micelles in aqueous solution and be attracted to
surfaces that are characterized by positive charge.
In the environment, the acidic form ionizes in water to a PFOA anion, and the ammonium
salt of PFOA rapidly dissociates. Physical and chemical properties and other reference
information for PFOA are provided in Table 2-1. These properties help to define the behavior of
PFOA in living systems and the environment. PFOA is a highly stable compound. It is a solid at
room temperature with a low vapor pressure. The melting point for PFOA is identified as 50 to
60 degrees Celsius (°C); vapor pressures increase at temperatures near the melting point.
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) - May 2016 15
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Table 2-1. Chemical and Physical Properties of PFOA
Property
Chemical Abstracts Service
Registry No. (CASRN3)
Chemical Abstracts Index
Name
Synonyms
Chemical Formula
Molecular Weight (g/mol)
Color/Physical State
Boiling Point
Melting Point
Vapor Pressure
Henry's Law Constant
pKa
KOC
P^OW
Solubility in Water
Half-life in Water (25°C)
Half -life in Air
Perfluorooctanoic Acid
335-67-1
2,2,3,3,4,4,5,5,6,6,7,7,8,8,8-
pentadecafluorooctanoic acid
PFOA; Pentadecafluoro-1-octanoic acid;
Pentadecafluoro-n-octanoic acid;
Octanoic acid, pentadecafluoro-;
Perfluorocaprylic acid;
Pentadecafluorooctanoic acid;
Perfluoroheptanecarboxylic acid;
C8HFi5O2
414.09
White powder (ammonia salt)
192.4°C; Stable when bound
54.3 °C
0.525 mm Hg at 25 °C (measured)
0.962 mm Hg at 59.25 °C (measured)
Not measureable
2.80
2.06
Not measurable
9.50 x 103 mg/L at 25 °C (estimated)
Stable
Stable when bound
Source
HSDB (2012); Lide (2007); SRC (2016)
HSDB (2012); Lewis (2004)
HSDB (2012); Lide (2007); SRC (2016)
HSDB (2012); Lide (2007); SRC (2016)
Hekster et al. (2003); HSDB (2012);
SRC (20 16)
ATSDR (2015); Kaiser et al. (2005)
ATSDR(2015)
SRC (20 16)
Higgins and Luthy (2006)
ATSDR (2015); EFSA (2008)
ATSDR (2015); Hekster et al. (2003);
HSDB (2012); Kauck and Diesslin
(1951); SRC (2016)
UNEP(2015)
UNEP(2015)
Notes'.
Kow = octanol-water partition co-efficient; K0c = organic carbon-water partitioning coefficient; g/mol = grams per mole
afhe CASRN given is for linear PFOA, but the toxicity studies are based on a mixture of linear and branched; thus, the RfD
applies to the total linear and branched.
PFOA is a strong acid that is generally present in solution as perfluorooctanoate anion. It is
water soluble and mobile in water, with an estimated log Koc of 2.06. PFOA is stable in
environmental media because it is resistant to environmental degradation processes, such as
biodegradation, photolysis, and hydrolysis. In water, no natural degradation has been
demonstrated, and dissipation is by advection, dispersion, and sorption to particulate matter.
PFOA has low volatility in ionized form, but can adsorb to particles and be deposited on the
ground and into water bodies. Because of its persistence, it can be transported long distances in
air or water, as evidenced by detections of PFOA in the arctic media and biota, including in polar
bears, ocean-going birds, and fish found in remote areas (Lindstrom et al. 201 la; Smithwick et
al. 2006). PFOA is present in ambient air and seawater globally (Ahrens et al. 2011; McMurdo et
al. 2008; Yamashita et al. 2005; Young et al. 2007).
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2.2 Occurrence and Sources of Exposure
PFOA and other PFASs have been discharged into the environment during use as processing
aids for fluoropolymers, by degradation of precursors, including fluorotelomer-based polymers,
and throughout the life cycle of products containing these compounds (i.e., from the point of
product manufacture through its use and disposal) (Washington et al. 2009, 2015a, 2015b).
PFOA and other PFASs are man-made chemicals, but because of their widespread use and
chemical and physical properties (persistence and mobility) they have been transported into
ground water, surface waters (fresh, estuarine, and marine), and soils in the vicinity of their
original source and at great distances. Point sources can result in significant exposure to people
in some areas. Major sources of PFOA are described below.
2.2.1 Surface Water and Ground Water
Water resources (i.e., surface water and ground water) are susceptible to contamination by
PFOA released from manufacturing sites, industrial use, fire/crash training areas, and industrial
or municipal waste sites where products are disposed of or applied. PFOA and other PFASs have
been reported in wastewater and biosolids as a result of manufacturing activities, disposal of
coated paper and other consumer products, and from washing stain-repellant fabrics (Renner
2009). Historically, land application of biosolids has been a source of PFOA and other PFASs in
surface water or ground water (Lindstrom et al. 201 Ib; Washington et al. 2010a, 2010b). The
phase-out of the use of these compounds in the United States is expected to reduce PFASs in
biosolids.
Some aqueous film forming foams used to combat aviation (or other hydrocarbon) fires
release PFOA to the environment (Seow 2013; USEPA 2014b). Surface and ground water
resources in close proximity to airports or other areas where these foams have been used can be
contaminated (see Moody et al. 2002). PFOA was reported at concentrations as high as 105 jig/L
in ground water near a concrete pad formerly used for military fire-training operations in
Michigan (ATSDR 2005; Moody et al. 2003). Surface water concentrations as a result of a
release of approximately 22,000 L of AFFF at L.B. Pearson International Airport in Toronto,
Canada, resulted in peak PFOA concentrations of 11.3 jig/L at the confluence of Etobicoke
Creek and Lake Ontario (Moody et al. 2002).
PFOA is not included as an analyte in the U.S. Geological Survey (USGS) National Water
Quality Assessment Program, and it is not monitored in water as part of EPA's National Aquatic
Resource Surveys. PFOA has been reported in U.S. water bodies, including the Tennessee River
(< 25-598 nanograms per liter [ng/L]), Mississippi River (< 1.0-125 ng/L), Lake Erie (21-47
ng/L), Lake Ontario (15-70 ng/L), and the Conasuaga River (253-1,150 ng/L) and Altahama
River (3.0-3.1 ng/L) watersheds in Georgia (Boulanger et al. 2004; Hansen et al. 2002; Konwick
et al. 2008; Nakayama et al. 2010). In another study, the USGS collaborated with the University
of Maryland and sampled three rivers and streams receiving effluent from 11 wastewater
treatment facilities in the Chesapeake Bay watershed (USGS 2011); samples were collected in
July and August 2010 from the Potomac River, the Patuxent River, and Saint Mary's Run. PFOA
concentrations ranged from 3.6 to 20 ng/L in the Patuxent River; from 7.5 to 12 ng/L in the
Potomac River; and from <2.0 to 47 ng/L in Saint Mary's Run.
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Studies show that PFOA occurs in marine waters. Yamashita et al. (2005) analyzed samples
from the Pacific Ocean, South China Sea, and Mid-Atlantic Ocean, as well as samples from
coastal waters of several Asian countries. PFOA was found at levels ranging from several
thousand picograms per liter (pg/L) in water samples collected from coastal areas in Japan to
tens of pg/L in the central Pacific Ocean. Yamashita et al. (2005) reported that PFOA was the
predominant PFAS detected in oceanic waters, followed by PFOS.
2.2.2 Drinking Water
Under EPA's UCMR 3, PFOA was monitored by approximately 5,000 PWSs (all PWSs
serving > 10,000 people, and a representative sample of 800 small PWSs) from 2013 through
December 2015. The minimum reporting level (MRL) for PFOA in this survey was 0.02 ug/L.
To-date, results for more than 36,000 samples have been reported by more than 4,800 PWSs for
PFOA. The remainder of the results are expected to be reported by mid-2016. PFOA was
measured at or above the MRL by approximately 2% of the PWSs. PFOA was reported above
0.07 ug/L by approximately 0.3% of PWSs that have reported results. Approximately 1% of
PWSs have reported data for which combined PFOA and PFOS results are above 0.07 ug/L.
The Environmental Working Group's (EWG3) National Drinking Water Database includes
PFOA analysis at 24 systems between 2004 and 2009 (EWG 2015). EWG obtained data
primarily from state drinking water offices; the database includes data from 47,677 water
systems in 45 states and the District of Columbia. The database showed that 24 systems reported
analyzing for PFOA; of these, five systems in Minnesota reported finding detectable levels. Four
of the systems had average concentrations below 0.01 ug/L. One system had an average
concentration of 0.09 ug/L and a maximum reported concentration of 0.25 ug/L.
PFOA detections in source water and drinking water are reported in several published
studies. These studies frequently involve targeted local sampling; thus, the findings are not
representative of national occurrence. For example, in New Jersey, monitoring of raw and
finished water between 2006 and 2008 revealed concentrations as high as 0.14 ug/L in finished
drinking water (NJDEP 2007; Post et al. 2009). In another study, PFOA concentrations in Little
Hocking, Ohio, ranged from 1.5 and 7.2 ug/L in the municipal water distribution system and up
to 14 ug/L in private wells between 2002 and 2005 (Emmett et al. 2006). A study in Minnesota
reported PFOA concentrations up to 0.9 ug/L in municipal, noncommunity, and private wells
between 2004 and 2008 (Goeden and Kelly 2006).
2.2.3 Food
PFOA ingestion from food is an important exposure source. PFOA was detected in a variety
of food products including snack foods, vegetables, meat, dairy products, human breast milk, and
fish, using data from Europe and North America as reported by Trudel et al. (2008). In North
America, snack foods, fish and shellfish, and potatoes were the food items estimated to
contribute the most to PFOA exposure, under intermediate and high-exposure conditions. In a
survey that included multiple food types, PFOA was the second-most frequently detected PFAS
3 For more information see http://www.ewg.org.
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and was present at high concentrations relative to other related compounds (Hlouskova et al.
2013). In a 2011 assessment of exposure to Americans, Lorber and Egeghy (2011) concluded
that food ingestion appears to be the primary route of exposure in adults, and dust and dietary
ingestion is the major contributor for young children, under typical exposure conditions. Recent
evidence shows that PFOA levels in food have been declining (Johansson et al. 2014).
Schecter et al. (2010) collected 10 samples of 31 commonly consumed foods from five
grocery stores in Dallas, Texas, in 2008 and analyzed them for PFOA. Equal weights of each
sample were combined and composited for analysis. Dietary intakes were estimated using data
from the 2007 U.S. Department of Agriculture food availability data set. For concentrations
below the limit of detection, a value of zero was assigned. The estimated per capita daily
estimate for exposure to PFOA was 60 nanograms per day (ng/day), or about 0.75 ng/day for an
80 kilogram (kg) adult. Based on a graphic presentation in the published paper, meat products
(n = 8) accounted for about 40 ng/day with the remaining 30% equally distributed between fish
(n = 7), vegetables (n = 7: three fat [olive oil, canola, margarine], one cereal, one apple, one
potato, and one peanut butter sample), and dairy and egg products (n = 9).
Tittlemier et al. (2007) conducted a Canadian total diet study that collected and analyzed 54
composite food samples. Samples were collected from 1992 to 2004 and represented fish and
seafood, meat, poultry, frozen entrees, fast food, and microwave popcorn. PFASs were detected
in nine composites (four meat, three fish and shellfish, one fast food, and one microwave
popcorn). PFOA and PFOS were most frequently found. The authors concluded that diet
represented approximately 60% of total PFASs exposure. PFOA was detected in roast beef,
pizza, and microwave popcorn at 0.74 to 3.6 ng/g, wet weight. The average daily PFOA
exposure was estimated at 70 ng.
Several studies are available from countries in Western Europe with diets that are
comparable to those in the United States. Fromme et al. (2007) collected duplicate diets for
15 male and 16 female healthy subjects (16 to 45 years old) in Germany. The median daily
dietary intake for PFOA was 2.9 nanograms per kilogram of body weight (ng/kg) (232 ng/day for
an 80 kg adult), with a 90th percentile intake of 672 ng/day. Haug et al. (2010) estimated
exposures in Norway using a market basket approach comprised of 21 foods, three drinking
water samples, one milk sample, and one tea sample. Total PFOA intake was estimated as
31 ng/day for the general Norwegian population. The highest levels were found in coffee, tea and
cocoa (2.1 ng/day), root vegetables/potatoes (0.66 ng/day), tap water (0.54 ng/day), soft drinks
(0.45 ng/day), and eggs (0.49 ng/day). Noorlander et al. (2011) estimated mean long-term daily
intakes of 0.2 ng/kg (16 ng/day for an 80 kg adult) in the Netherlands using a pooled composite
from foods purchased in retail chains with nationwide coverage; the 99th percentile value was
0.5 ng/kg (40 ng/day). Important sources were vegetables, fruits, and flour.
Human studies have shown that PFOA is transferred from mother to infant via cord blood
and breast milk. A recent study showed that breast milk contributed > 83% of the PFOA
exposure in 6-month-old infants (Haug et al. 2011). Additional information on concentrations of
PFOA in breast milk is provided in section 2.5.
PFOA has been detected in beef as a result of cattle ingesting contaminated feed. When cattle
were exposed to feed contaminated with 13 ng PFOA/kg wet weight, PFOA accumulated in the
liver (9 ng/kg) and in muscle (7 ng/kg) (Vestergren et al. 2013). The study also detected PFOA
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) - May 2016 19
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in cow's milk at 6.7 ng/L. In addition, evidence suggests that livestock accumulate PFOA by
grazing in fields where biosolids were applied (Renner 2009; Vestergren et al. 2013).
Bioaccumulation in fish and other edible aquatic organisms is another route for potential
dietary exposures (Bhavsar et al. 2014; Renzi et al. 2013; Stahl et al. 2014). EPA analyzed fish
fillet tissue samples from U.S. rivers and from the Great Lakes as part of EPA's National
Aquatic Resource Surveys. These analyses included characterizing perfluorinated compounds
(PFCs) in freshwater fish on a national scale during EPA's 2008-2009 National Rivers and
Streams Assessment, and on a regional scale during the Great Lakes Human Health Fish Tissue
Study component of the EPA 2010 National Coastal Condition Assessment. Fish were collected
from randomly selected locations, including 162 urban river sites and 157 nearshore Great Lake
sites, and analyzed for 13 PFASs. Results showed that 80% of urban river fish samples and
100% of Great Lakes fish samples contained some detectable PFASs. PFOS was the most
frequently detected chemical (in 73% of river fish samples and 100% of Great Lakes fish
samples). PFOA was not detected in river fish fillet samples, but it was detected in 12% of the
Great Lakes samples. In the 2010 Great Lakes sampling, PFOA was detected in 19 out of 157
samples at a maximum concentration of 0.97 ng/g. The differences in PFOA detections between
river and Great Lakes fish samples could be due to the availability of a more sensitive PFAS
analytical method with lower detection limits when the Great Lakes study was initiated. Cooking
offish does not reduce the levels of PFOA in the fish (or the consumer's dietary exposure)
(Bhavsar etal. 2014).
PFOA has been detected in wild caught and farmed fish, presumably because of
bioaccumulation and/or trophic transfer. Bhavsar et al. (2014) found that PFOA concentrations
were higher in wild-caught fish than farmed fish, and suggested that fish caught near
contaminated sites could represent an important exposure source among recreational and
subsistence fishers.
In a survey of French adult freshwater anglers, PFOA was a major contributor of total PFAS
exposure from fish. Although some individuals had higher exposures, overall values for this
population were close to those for the general population (Denys et al. 2014). In a study of
French adults who consumed large amounts of seafood (n = 993), mean lower bound exposure to
PFOA was 1.16 ng/kg/day (92.8 ng/day for an 80 kg adult) compared to a lower bound of none
in the general population (n = 1918). The mean upper bound values were 2.06 and 0.74
ng/kg/day (164.5 to 59.2 ng/day), respectively, for the same highly exposed and general
population groups (Yamada et al. 2014). In a sub-study that was restricted to 106 pregnant
women, the upper bound mean was 1.52 ng/kg/day (121.6 ng/day) and the 95th percentile upper
bound was 2.41 ng/kg/bw/day (192.8 ng/day).
PFOA can occur in plants grown in soils containing PFOA. For example, PFOA was taken
up by corn when grown in biosolid-amended soil; however, the chemical remained in the roots
and did not accumulate in edible parts of the plant (Krippner et al. 2014). PFOA accumulation in
fruit crops tends to be lower than in shoot or root crops, presumably because there are more
compartments through which PFOA would have to pass to reach the edible portion of the plant
(Elaine etal. 2014).
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PFOA was previously used in the manufacturing of several types of food packaging; in
January 2016, the U.S. Food and Drug Administration (FDA) amended its food additive
regulations to no longer allow for the use of perfluoroalkyl ethyl-containing food-contact
substances as oil and water repellants for paper and paperboard that comes in contact with
aqueous and fatty foods (81 FR 5). PFOA is a breakdown product of the perfluorooctylethanol
telomer alcohol used to make coatings for or additives in food contact paper where it adds a
moisture or oil barrier to paper-type packaging, including microwave popcorn bags, fast food
wrappers, candy wrappers, and pizza box liners (Begley et al. 2005). When used in this way,
PFOA can migrate into foods from the packaging material. In a study conducted by FDA, Begley
et al. (2005) was able to extract (4 micrograms per square decimeter [jig/dm2] paper) of PFOA
into food oil before cooking and another 7 jig/dm2 from paper after cooking. Based on these
results, Begley et al. (2005) concluded that paper with treated coatings had a high potential for
migration of fluorochemical to food.
Food can become contaminated with PFOA from preparation in nonstick cookware coated
with polytetrafluoroethylene (PTFE) (Teflon™). PFOA is a processing aid in the manufacture of
PTFE. Begley et al. (2005) also evaluated migration of PFOA to foods from cooking in
Teflon™-lined cookware and found it to be much lower (0.03 jig/dm2 polymer) than migration
from coated paper. In this study, new pans leached more compared to those that had been used
before.
2.2.4 Ambient Air
A number of PFASs are precursors of PFOA and degrade to PFOA in the environment via
biotic and abiotic degradation. Some of these precursors are volatile and contribute to the
formation of airborne PFOA. Indoor air sampling reportedly contains higher concentrations of
these precursors than outdoor air (Vierke et al. 2012). Langer et al. (2010) reported detections of
PFOA and precursors in indoor air samples from home residences and at stores that sold outdoor
equipment, furniture, and carpet. Fraser et al. (2013) found that PFOA in serum was significantly
correlated with air levels collected in offices, likely associated with carpeting, furniture, and
paint.
PFOA can be emitted from nonstick cookware coated with PTFE. Schlummer et al. (2015)
found that at typical cooking temperatures (< 230°C), perfluoroalkylcarboxylic acids (C4 to C12)
dominated (4.75 ng per hour) by PFOA and perfluorobutanoic acid (PFBA) were released to the
atmosphere; when pans were overheated PFBA and perfluoro-n-pentanoic acid (PFPeA) were
dominant (> 260°C). Emissions were far greater at higher temperatures (12,190 ng per hour at
370 °C; Schlummer et al. 2015). Emissions are expected to decline with use of the product. The
authors hypothesized that most of the emissions would end up in household dusts.
Based on its environmental fate properties, PFOA has low volatility. However, PFOA has
been reported in ambient air, largely bound to particulate matter. It can be transported long
distances via the atmosphere and has been detected at low concentrations in areas as remote as
the Arctic (Shoeib et al. 2006) and Antarctic (Del Vento et al. 2012). PFOA levels in outdoor air
were measured in a variety of locations, most of which are countries outside the United States.
Fromme et al. (2009) reported mean levels of 2 picograms per cubic meter (pg/m3) in particulate
matter for eight samples collected in the summer in Albany, New York with a mean of 3.2 pg/m3
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) - May 2016 21
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present in the gas phase. Mean air concentrations in Spain and England were 6.1 pg/m3 and
3.5 pg/m3, respectively (Beser et al. 2011; Goosey and Harrad 2012). In a study conducted in
China, airborne PFOA concentrations were similar (Liu et al. 2015). Areas near wastewater
treatment plants, waste incinerators, and landfills can be point sources for PFOA in outdoor air
(Ahrens et al. 2011). PFOA-derived telomer alcohols can also be present in air (Jogsten et al.
2012).
2.2.5 Indoor Dust
Because of its widespread use in carpets, upholstered furniture, and other textiles, PFOA has
been detected in indoor dust from homes, offices, vehicles, and other indoor spaces. Although
some of these uses have been phased out, exposure could continue in legacy products and
imported goods. As reported by Fraser et al. (2013), particulate matter from fabrics and carpeting
are believed to be the source for the PFOA-containing dusts found in homes, offices, and
automobiles.
A 2013 survey (Fraser et al. 2013) detected PFOA in samples of house dust (23.7 ng/g),
office dust (32.0 ng/g), and vehicles (11.4 ng/g) collected at sites by 31 participants in Boston,
Massachusetts. The Wisconsin Department of Health and Human Services collected vacuum
cleaner contents from 39 homes as a means of evaluating the concentration of PFOA and 15
other PFASs in dust (Knobeloch et al. 2012). The median PFOA concentration was 44 ng/g.
PFOA, PFOS and perfluorohexane sulfonate (PFHxS) accounted for about 70% of the total
PFASs present in the dust. Lorber and Egeghy (2011) assessed Americans' PFOA exposure and
concluded that ingestion of household dust and food are primary routes of PFOA exposure for
2-year-old children. For median exposed children, exposures were estimated to be 13 and 8 ng/d
from dust and food, respectively. For highly exposed children (at the 95th percentile), PFOA
exposure from dust was estimated to be three times that from food.
Jogsten et al. (2012) collected dust samples from 10 selected homes in Catalonia, Spain, and
analyzed them for 20 PFASs. All samples contained PFOA; the levels ranged from 1.5 to
13.9 ng/g. An important outcome of this study was the identification of PFOA volatile telomer
alcohol derivatives in the dust samples at concentrations of up to 1.3 ng/g. The 8:2 telomer
alcohols degrade metabolically to PFOA once ingested. A study conducted in Belgium also
found that PFOA was present in home (median: 0.7 ng/g dry weight) and office dust (median:
2.2 ng/g dry weight) (D'Hollander et al. 2010). The highest of the indoor dust concentrations of
those sampled (114 ng/g) were found in homes in Germany (Xu et al. 2013).
2.2.6 Soils
PFOA persists in soils near manufacturing facilities and disposal sites (Xiao et al. 2015) and
in areas, such as military bases, where firefighting foams containing PFOA were heavily used
(Filipovic et al. 2015). Measured concentrations of PFOA in surface soils range from 8.0 ng/g
(Xiao et al. 2015) to 287 ng/g (Filipovic et al. 2015). These studies focused on two sites, the first
in the Minneapolis-St. Paul, Minnesota, metropolitan area where PFASs were manufactured and
disposed of, and the second on a former military airport in Sweden abandoned in 1994, where
firefighting foams containing PFOA had been used. In both cases, there was ground water
contamination. Xiao et al. (2015) determined that levels of PFOA in soils increased with depth,
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) - May 2016 22
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providing evidence for migration into ground water (see also section 2.2.1). Filipovic et al.
(2015) found that PFOA concentrations in soil cores remained high more than 30 years after
usage was discontinued.
Incidental ingestion of soils represents a potential exposure route for PFOA. Regional and
geographic differences in soil characteristics can influence PFOA concentrations. Soil
contamination tends to occur at manufacturing sites of producers and users, where disposal of
treated products has occurred (i.e., landfills), and potentially where biosolids containing PFASs
are applied. Calculated residence time in soils suggests that persistence in the environment will
extend well beyond the time that PFOA manufacturing ends (Zareitalabad et al. 2013).
Contaminated soils also can be transported offsite via water and wind.
2.2.7 Biosolids
Biosolids are sometimes applied as an amendment to soils as fertilizers; in some cases, the
biosolids can contain PFOA. For example, in May 2007, a Decatur, Alabama, manufacturer that
used PFASs notified the Decatur Utilities Dry Creek Waste Water Treatment plant that it had
unknowingly discharged large amounts of perfluorocarboxylic acid precursors (PFOA and
perfluorododecanoic acid [PFDA]) to the utility (USEPA 201 la). The Decatur treatment
plant also received wastewater from several other industries in the area that manufactured or
used a variety of PFAS-containing materials. The incident was reported to EPA and other
government agencies because biosolids from the wastewater plant had been applied to 5,000
acres of privately owned agricultural fields for the previous 12 years (1996 to 2008).
Testing revealed that the biosolids from the Decatur plant contained PFOA, PFOS and other
PFASs. Concentrations in nine soil samples from the area ranged from 589 to 1,296 parts per
billion (ppb) PFOA and 55 to 2,531 ppb PFOS. Subsequently, private wells, ponds, and other
surface waters near the biosolids application sites were sampled and found to contain PFOS and
PFOA, in some cases at levels greater than EPA's provisional HA values. Several additional
rounds of sample collection from the impacted areas confirmed the presence of PFASs, including
PFOA and PFOS in the media tested (Lindstrom et al. 201 Ib; USEPA 201 la; Washington et al.
2010a, 201 Ob).
PFASs were not analyzed in the 2004 EPA Total National Sewage Sludge Survey (TNSSS),
as analytical methods were not available when analytes were selected. Venkatesan and Halden
(2013) re-analyzed archived samples for PFASs from the TNSSS in five composite samples,
which represented 94 wastewater treatment facilities from 32 U.S. states and the District of
Columbia in 2001. PFOS was the most abundant PFC identified (mean 403± 127 |ig/kg dry
weight), followed by PFOA (mean 34 ± 22 |ig/kg dry weight). Armstrong et al. (2016) collected
biosolid samples every 2 months from a large municipal water recovery facility between 2005
and 2013. The highest mean PFOA concentration reported was 23.5 |ig/kg dry weight. Yoo et al.
(2011) found PFOA and PFOS in plants (fescue, barley, bluegrass, and Bermuda grass) grown in
soils amended with biosolids. Concentrations of PFOA ranged from 9.9 to 202.7 |ig/kg.
Concentrations in biosolids are expected to decline because of the phase-out of the use of PFOS
and PFOA in manufacturing and industrial processes.
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2.2.8 Consumer Products
Other materials that result in potential human exposure include legacy use and imported
goods or continuing uses. Some examples of these uses are listed below.
• Stain/water repellants on clothing, bedding materials, upholstered furniture, carpets, and
automobile interiors (e.g., Stainmaster™, Zonyl™, Nuva™, Unidyne™, Baygard™)
(Walters and Santillo 2006); these materials can be a particularly important exposure
route for infants and children because of their hand-to-mouth behaviors.
• Cooking surfaces (e.g., Teflon™)
• Toothpaste, shampoos, cosmetics
• Polishes and waxes
• Electronics
• Flame repellants
• Paints, varnishes, sealants
• Lubricants/surfactants/emulsifiers (continuing use)
• Food containers and contact paper4
• Pesticide
• Aqueous film forming foams (continuing use; used for firefighting)
• Electronics
• Textiles (e.g., Gore-Tex™) and leather
• Plumbing tape
• Cleaning products
2.3 Environmental Fate
2.3.1 Mobility
PFOA is water soluble and has been found in surface water, ground water, and drinking
water. It has low volatility in ionized form, but can adsorb to particles in air; because of its
persistence, it can be transported long distances to the Arctic (Shoeib et al. 2006) and Antarctic
(Del Vento et al. 2012). PFOA has a log Koc of 2.06 and does not easily adsorb to sediments or
aquifer materials; therefore, it tends to stay in the water column.
2.3.2 Persistence
PFOA is stable in the environment and resistant to hydrolysis, photolysis, volatilization, and
biodegradation (see Table 2-1). No biodegradation or abiotic degradation processes have been
found; the only dissipation mechanisms in water are dilution, advection, and sorption. Yamada et
4 PFOA was used in some grease-proofing paper coatings or additives that can contribute to its presence in foods
(Begley et al. 2005). However, in January 2016, FDA amended their food additive regulations to no longer allow for
the use of perfluoroalkyl ethyl containing food-contact substances as oil and water repellants for paper and
paperboard for use in contact with aqueous and fatty foods (81 FR 5).
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al. (2005) determined that typical municipal waste incinerators destroy PFOA on textiles and
paper and do not release it into the atmosphere.
2.3.3 Bioaccumulation
Several criteria can be used to assess bioaccumulation, including octanol-water partition
coefficient (Kow), bioconcentration factors (BCF), bioaccumulation factors (BAFs), and
biomagnification or trophic magnification factors (BMFs or TMFs, respectively) (Gobas et al.
2009). The Kow and BCF metrics are typically based on partitioning of organic chemicals into
octanol or lipids of biota. For PFOA, partitioning appears to be more related to protein binding
properties than its lipid partitioning. Thus, the Kow is not a reliable measure of bioaccumulation
potential for PFOA (EFSA 2008; UNEP 2015). Information from field studies, BCFs, BMFs,
and TMFs provide the most conclusive evidence of accumulation of chemicals in food webs
(Gobas et al. 2009) and are the more appropriate metrics for gauging the potential for
accumulation of PFOA in fish, wildlife, and humans.
Because of the physical-chemical properties of PFOA, Kow cannot be reliably measured
(Table 2-1; UNEP 2015; USEPA 2014b). Model estimates of Kow have been reported; however,
verification that these chemicals are within the domain of the models is often not provided.
Therefore, the validity of the use of such models is questionable (EFSA 2008; UNEP 2015),
Available BCFs determined from lab studies have been reported and generally fall below
traditional criteria used to assess bioaccumulation (e.g., Martin et al. 2003c). It is recognized,
however, that BCFs determined by existing standard methods derived from lipid-partitioning are
not an appropriate metric for assessing bioconcentration of PFOA (EFSA 2008; UNEP 2015).
Although evidence of PFOA accumulation in many organisms has been documented, reported
BAFs and BCFs for the chemical also fall below traditional criteria used to assess
bioaccumulation potential (Loi et al. 2011; Martin et al. 2003a, 2003b; Morikawa et al. 2005;
Quinete et al. 2009).
Field evidence of PFOA biomagnification, considered to be the preferable metric for
assessing bioaccumulation potential (Gobas et al. 2009), has been documented in many
organisms from many locations worldwide (UNEP 2015). Trophic magnification has also been
evaluated (Environment Canada and Health Canada, 2012; Houde et al. 2006; Kelly et al. 2009;
Loi et al. 2011; Martin et al. 2004). Some field trophic studies revealed TMFs greater than 1,
which indicates that PFOA accumulated and increased in concentration with increasing trophic
level; other studies reported TMFs less than 1 for some food webs. The weight of evidence for
trophic magnification was deemed sufficient to consider PFOA to be bioaccumulative by the
Stockholm Convention Persistent Organic Pollutants Review Committee (UNEP 2015).
2.4 Toxicokinetics
Uptake and egress of PFOA from cells is largely regulated by transporters in cell membranes
(Anzai et al. 2006; Cheng et al. 2006; Klaassen and Aleksunes 2010; Nakagawa et al. 2007,
2009; Weaver et al. 2009, 2010; Yang et al. 2010). PFOA is absorbed from the gastrointestinal
tract as indicated by the serum measurements in humans and treated animals. In serum, it is
electrostatically bound to albumin, occupying nine to 12 sites, and sometimes displaces other
substances such as nutrients and pharmaceuticals that normally would occupy a site (MacManus-
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Spencer et al. 2009; Salvalaglio et al. 2010; Wu et al. 2009a). Linear PFOA chains display
stronger binding than branched chains (Beesoon and Martin 2015). Binding causes a change in
the conformation of serum albumin, thereby changing its affinity for the endogenous compounds
it normally transports.
PFOA is distributed to tissues by a process requiring transporters. Accordingly, the tissue
levels vary from organ to organ as demonstrated by Kemper (2003). The highest tissue
concentrations are usually those in the liver. Liver accumulation in males is greater than that in
females. Other tissues with a tendency to accumulate PFOA are the kidneys, lungs, heart, and
muscle, plus the testes in males and uterus in females (Kemper, 2003). PFOA is not metabolized,
thus any effects observed in laboratory toxicological studies are the result of parent compound,
not metabolites.
Electrostatic interactions with proteins are an important toxicokinetic feature of PFOA.
Studies demonstrate binding or interactions with receptors (e.g., peroxisome proliferator
activated receptor alpha [PPARa], triiodothyronine [T3]), transport proteins and enzymes
(Luebker et al. 2002; Weiss et al. 2009; L. Zhang et al. 2013). Saturable renal resorption of
PFOA from the glomerular filtrate via transporters in the kidney tubules is a major contributor to
the long half-life of this compound in humans (Nakagawa et al. 2007, 2009; Weaver et al. 2010;
Yang et al. 2009, 2010). Branched-chain PFOAs are less likely to be resorbed than the linear
molecules based on half-life information in humans (Y. Zhang et al. 2013). All toxicokinetic
models for PFOA are built on the concept of saturable renal resorption first proposed by
Andersen et al. (2006). Some PFOA is removed from the body with bile (Genuis et al. 2010), a
process that also is transporter-dependent. Accordingly, the levels in fecal matter represent both
unabsorbed material and material that is discharged to the intestines with bile.
During pregnancy, PFOA is present in the placenta and amniotic fluid in both animals
(Fenton et al. 2009; Hinderliter et al. 2005) and humans (T. Zhang et al. 2013). Post-delivery,
PFOA is transferred to offspring through lactation in a dose-related manner (Hinderliter et al.
2005, Fenton et al. 2009). Maternal serum levels decline as those in the pups increase. This also
occurs in humans as demonstrated in the study by Mondal et al. (2014) of breastfeeding women
and their infants in West Virginia and Ohio.
The half-life in humans for occupationally exposed workers (Olsen et al. 2007) was 3.8 years
(95% CI [1.5, 9.1]). Bartell et al. (2010) determined an average half-life of 2.3 years based on a
study of the decreases in human serum levels after treatment of drinking water for PFOA
removal was instituted by the Lubeck Public Services District in Wqest Virginia and the Little
Hocking Water Association in Ohio. This is the value used for humans in this assessment
because it applies to the general population and reflects humans whose exposure came primarily
from their PWS. Half-lives are reported to be shorter in animals than for humans: 21 days
(females) and 30 days (males) for monkeys (Butenhoff et al. 2004b); 11.5 days (males) and 3.4
hours (females) for Sprague-Dawley rats (Kemper 2003); 27.1 days (male) and 15.6 days
(female) for CD-I mice (Lau et al. 2006). Although the animal half-lives are shorter than
humans, so, too, are their average lifetimes. In early life, the half-lives are nearly the same for
both genders, but once the animals reach sexual maturity resorption increases in male rats,
prolonging the half-lives (Hinderliter et al. 2006; Hundley et al. 2006). This change appears to be
under the control of hormones in both males and females (Cheng et al. 2006; Kudo et al. 2002).
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2.5 Human Biomonitoring Data
The Fourth National Report on Human Exposure to Environmental Chemicals from the
Centers for Disease Control and Prevention (CDC 2009) included exposure data for PFOA from
2003 to 2004 collected by the National Health and Nutrition Examination Survey (NHANES).
PFOA was detected in 99.7% of the general U.S. population. Since that time, CDC has issued
several updates, the most recent of which was released in 2015 (CDC 2015). Taken together, the
data suggest that PFOA concentrations in human serum in the U.S. generally declined between
1999 and 2012. The geometric mean PFOA concentration in human serum decreased from 5.2 to
2.1 ug/L, and the 95th percentile concentration decreased from 11.9 to 5.7 ug/L. During this time,
there has been a major reduction in environmental emissions by the manufacturers as well as a
phase-out of production of C-8 compounds in the United States. Analysis of the NHANES 2003-
2004 subsample demonstrated higher levels of PFOA and PFOS in males and a slight increase in
levels of PFOS with age (Calafat et al. 2007a, 2007b).
Precursors might also form PFOA in the body; this represent an important uncertainty in
characterizing exposure as measured by blood serum. For example, Lorber and Egeghy (2011),
indicated that the precursor fluorotelomer alcohols (FTOHs) and polyfluoroalkyl phosphoric
acids (PAPs) would add to exposure but there is uncertainty as to the magnitude of the effect.
The authors concluded that precursors "could very well contribute half or more of what is
eventually measured as PFOA in the blood."
Evidence shows that PFOA is distributed within the body and can be transferred from
pregnant women to their unborn children and offspring. T. Zhang et al. (2013) collected serum
and cord blood samples from 30 pregnant women in China. The maternal blood contained
variable levels of 10 PFASs, eight acids, and two sulfonates. The mean maternal blood
concentration for PFOA was 3.35 nanograms per milliliter (ng/mL). The mean was greater than
the median, indicating a distribution skewed toward the higher concentrations. Compared to the
mean PFOA blood levels in the pregnant women, the mean levels in cord blood (1.95 milligrams
per milliliter [mg/mL]) was 47% of that in the mother's blood.
PFOA has been detected in breast milk (Tao et al. 2008; Volkel et al. 2008) and cord blood
(Apelberg et al. 2007; Monroy et al. 2008) at concentrations above the limit of quantification.
Mondal et al. (2014) evaluated serum samples from breastfeeding women and their infants in
West Virginia and Ohio. For each month of breastfeeding, maternal serum levels of PFOA were
reduced by 3% (95% CI: 2%-5%) and infant serum levels increased by 6% (95% CI: 1%-10%]).
A publication from the French total diet study (Cariou et al. 2015) also examined human breast
milk as an exposure route for infants using 61 mother-infant pairs. PFOA was detected in 77% of
the breast milk samples, with a mean concentration of 0.041 ng/mL and a maximum
concentration of 0.308 ng/mL. The regression coefficient for the association between the
maternal serum concentration and the detected breast milk concentrations was 0.72 (n = 10).
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3.0 PROBLEM FORMULATION
3.1 Conceptual Model
The conceptual model provides useful information to characterize and communicate the
potential health risks related to PFOA exposure from drinking water. The sources of PFOA, the
routes of exposure for biological receptors of concern (e.g., various human activities related to
ingested tap water such as drinking, food preparation, and consumption), and the potential
assessment endpoints (e.g., effects such as liver toxicity and developmental effects), and adverse
health effects in the populations at risk due to exposure to PFOA are depicted in the conceptual
diagram below (Figure 3-1).
STRESSOR
SOURCES
EXPOSURE ROUTE
RECEPTORS
ENDPOINTS
Showering/
Bathing
Incidental
Ingestion While
Showering
Cooking with Water
Incidental
Ingestion While
Showering
Pregnant and
actating Women
Cardiovascular/
Serum Lipid
Effects
Liver
Effects
Developmental
Effects
Reproductive
Effects
Aviation Film
Forming
Foams
Legend
Quantitatr
Assessmen
Qualitativ
Assessmen
Immune
Effects
Thyroid
Effects
Figure 3-1. Conceptual Model for PFOA in Finished Drinking Water
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3.1.1 Conceptual Model Diagram for Exposure via finished Drinking Water
The conceptual model is intended to explore potential links of exposure to a contaminant or
stressor with the adverse effects and toxicological endpoints important for management goals,
including the development of drinking water HA values. Boxes that are more darkly shaded
indicate pathways that were considered quantitatively in estimating the advisory level, whereas
the lightly shaded boxes were only considered from a qualitative perspective.
3.1.2 Factors Considered in the Conceptual Model for PFOA
Stressors: For this HA, the stressor is PFOA in drinking water from public water facilities or
private wells.
Sources: Sources of PFOA include both ground and surface waters used for drinking.
Multiple potentially important sources of PFOA and precursors exist in addition to drinking
water, such as foods, indoor dust in a home or work environment, indoor and outdoor air, soil,
consumer products within the homes or place of work including children's schools, and
industrial products. The relative contribution of drinking water versus other sources is addressed
in the Relative Source Contribution section of the document (section 3.2.5). This HA applies
only to drinking water.
Routes of exposure: Exposure to PFOA from contaminated drinking water sources can occur
via oral exposure (drinking water, cooking with water, and incidental ingestion from showering);
dermal exposure (contact of exposed parts of the body with water containing PFOA during
bathing or showering, dishwashing); and inhalation exposure (during bathing or showering or
using a humidifier or vaporizer). There is limited information identifying health effects from
inhalation or dermal exposures to PFOA in humans and animals. Therefore, these routes of
exposure are not quantitatively used in the derivation of the HA. PFOA has a low vapor pressure
and is not expected to be present in air except as bound to paniculate matter and in aerosols
formed from devices such as shower heads and humidifiers that aerosolize tap water. Toxicity
data are available for oral exposure from drinking water, but not the other exposure routes
(inhalation and dermal exposures). PFOA is not removed by heating water and can increase in
concentration when the water is boiled.
Receptors: The receptors are those in the general population (adults, infants and children)
who could be exposed to PFOA from tap water through dermal contact and inhalation and/or
ingestion at their homes, workplaces, schools, and daycare centers.
Endpoints: Epidemiology data report associations between PFOA exposure and high
cholesterol, increased liver enzymes, decreased vaccination response, thyroid disorders,
pregnancy-induced hypertension and preeclampsia, and cancer (testicular and kidney) (see
section 4.1.2). These studies provide varying levels of support for the effects associated with
PFOA exposure in the animals studies used for quantification of the HA. Cholesterol, liver
enzymes, and thyroid effects were examined in numerous studies in different populations, but the
pregnancy complications of hypertension and preeclampsia in women and testicular cancer in
young men were only studied in a high-exposure community (located in the vicinity of a PFOA
production plant in West Virginia; i.e., C8 Health Project). The C8 Science Panel assessed the
links between PFOA and several diseases and concluded that a probable link existed between
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PFOA and the observed kidney and testicular tumors among the population evaluated (see
section 4.1.2).
The associations for most epidemiology endpoints are mixed. Although mean serum values
are presented in the human studies, actual estimates of PFOA exposure (i.e., doses/duration) are
not currently available. Thus, the serum level at which the effects were first manifest and
whether the serum had achieved steady state at the point the effect occurred cannot be
determined. It is likely that some of the human exposures that contribute to serum PFOA values
come from PFOA derivatives or precursors that break down metabolically to PFOA. These
compounds could originate from PFOA in diet and materials used in the home, which creates
potential for confounding. In addition, most of the subjects of the epidemiology studies have
many PFASs and/or other contaminants in their blood. Although the study designs adjust for
other potential toxicants as confounding factors, their presence constitutes a level of uncertainty
that is usually absent in the animal studies.
Taken together, the weight of evidence for human studies supports the conclusion that PFOA
exposure is a human health hazard. At this time, EPA concludes that the human studies are
adequate for use qualitatively in the identification hazard and are supportive of the findings in
laboratory animals. EPA plans to begin another effort to determine the range of perfluoroalkyl
compounds for which an Integrated Risk Information System (IRIS) assessment is needed, as
indicated in the 2015 IRIS Multi Year Agenda.5
For PFOA, oral animal studies of short-term, subchronic, and chronic duration are available
in multiple species including monkeys, rats, and mice (see section 4.1.1). Adverse effects
observed following exposure to PFOA include liver toxicity (hypertrophy, necrosis, and effects
on the metabolism and deposition of dietary lipids), kidney toxicity, and developmental effects
(survival, body weight changes, reduced ossification, altered puberty, and retarded mammary
gland development), immune effects, and cancer. EPA quantitatively evaluated (i.e., modeled
serum concentrations) for the liver, developmental, immune, and cancer effects.
In most animal studies, changes in relative and/or absolute liver weight appears to be the
most common effect observed with or without other hepatic indicators of adversity identifying
increased liver weight as a common indicator of PFOA exposure. The liver also contains the
highest levels of PFOA when analyzed after test animal sacrifice. The increases in liver weight
and hypertrophy, however, also can be associated with activation of cellular PPARa receptors,
making it difficult to determine if this change is a reflection of PPARa activation or an indication
of PFOA toxicity. The PPARa response is greater in rodents than it is in humans. EPA evaluated
liver disease and liver function resulting from PFOA exposure in studies where liver weight
changes and other indicators of adversity such as necrosis, inflammation, fibrosis and/or steatosis
(fat accumulation in the liver) or increases in liver or serum enzymes indicative of liver damage
were observed. Only the doses associated with the adverse effects were used for the
quantification of risk.
5 For more information on the IRIS agenda see https://www.epa.gov/iris/iris-agenda.
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3.2 Analysis Plan
3.2.1 Health Advisory Guidelines
Assessment endpoints for HAs can be developed for both short-term (1-day and 10-day) and
lifetime exposure periods using information on the noncarcinogenic and carcinogenic
toxicological endpoints of concern. Where data are available, endpoints reflect susceptible and/or
more highly exposed populations.
• A 1-day HA is typically calculated for an infant (0 to 12 months or a 10-kg child),
assuming an acute exposure to the chemical; it is generally derived from a study of less
than 7 days duration.
• A 10-day HA is typically calculated for an infant (0 to 12 months or a 10-kg child),
assuming a limited period of exposure of 1 to 2 weeks; it is generally derived from a
study of 7 to 30 days duration.
• A lifetime HA is derived for an adult (> 21 years old or an 80-kg adult), and assumes an
exposure period over a lifetime (approximately 70 years). It is usually derived from a
chronic study of 2 years duration, but subchronic studies can be used by adjusting the
uncertainty factor employed in the calculation. For carcinogens, the HA documents
typically provide the concentrations in drinking water associated with a range of risks
(from one excess cancer case per 10,000 persons exposed to one excess cancer case per
million persons exposed) for Group A and B carcinogens and those classified as known
or likely carcinogens (USEPA 1986, 2005). Cancer risks are not provided for Group C
carcinogens, or those classified as "suggestive," unless the cancer risk has been
quantified.
3.2.2 Establishing the Data Set
The Health Effects Support Document for Perfluorooctanoic Acid (PFOA) (USEPA 2016a)
provides the health effects basis for development of the HA, including the science-based
decisions providing the basis for estimating the point of departure (POD). To develop the HESD
for PFOA, EPA assembled available information on toxicokinetics, acute, short-term, subchronic
and chronic toxicity along with developmental and reproductive toxicity, neurotoxicity,
immunotoxicity, genotoxicity and cancer in humans and animals. For a more detailed description
of the literature review search and strategy for inclusion and exclusion of studies, see the
Forward and Appendix A of the HESD for PFOA.
Briefly, through a literature search, studies were identified for retrieval, review, and inclusion
in the document using the following criteria:
• The data contribute substantially to the weight of evidence for any of the toxicity
endpoints.
• Elements of the study design merit its inclusion in the draft assessment based on its
contribution to the mode of action (MOA) or the quantification approach.
• The study elucidates the MOA for any toxicity endpoint or toxicokinetic property
associated with PFOA exposure.
• The effects observed differ from those in other studies with comparable protocols.
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• The study was relevant to drinking water exposures and to the U.S. population.
In addition, an evaluation of available data was performed by EPA to determine data
acceptability. The following study quality considerations from U.S. EPA's (2002)^4 Review of
the Reference Dose and Reference Concentration Processes were used in selection of the studies
for inclusion in the HESD and development of the HA.
• Clearly defines and states hypothesis.
• Adequately describes the study protocol, methods, and statistical analyses.
• Evaluates appropriate endpoints. Toxicity depends on the amount, duration, timing and
pattern of exposure, and could range from frank effects (e.g., mortality) to more subtle
biochemical, physiological, pathological or functional changes in multiple organs and
tissues.
• Applies appropriate statistical procedures to determine an effect.
• Establishes dose-response relationship (i.e., no observed adverse effect level (NOAEL)
and/or lowest observed adverse effect level (LOAEL) or data amenable to modeling of
the dose response to identify a POD for a change in the effect considered to be adverse
[out of the range of normal biological viability]. The NOAEL is the highest exposure
level at which there are no biologically significant increases in the frequency or severity
of adverse effects between the exposed population and its appropriate control. The
LOAEL is the lowest exposure level at which there are biologically significant increases
in frequency or severity of adverse effects between the exposed population and its
appropriate control group.
The studies included in the HESD and HA were determined to provide the most current and
comprehensive description of the toxicological properties of PFOA and the risk it poses to
humans exposed through their drinking water.
After the available, reliable studies were evaluated for inclusion in the HESD and HA,
critical studies were selected for consideration based on factors including exposure duration
(comparable to the duration of the HAs being derived), route of exposure (e.g., oral exposure via
drinking water, gavage, or diet), species sensitivity, comparison of the POD with other available
studies demonstrating an effect, and confidence in the study (USEPA 1999). Uncertainty factors
appropriate for the studies selected are then applied to the potential PODs to account for
variability and uncertainty in the available data.
3.2.3 Approach for HA Calculation
For PFOA, toxicity and exposure data were used to develop a lifetime HA. EPA used
measures of effect and estimates of exposure to derive the lifetime HA using the following three-
step process:
Step 1: Adopt a Reference Dose (RfD) or calculate an RfD using the appropriate point of
departure (POD). The RfD is an estimate (with uncertainty spanning perhaps an order of
magnitude) of a daily human exposure to the human population (including sensitive subgroups)
that is likely to be without an appreciable risk of deleterious effects during a lifetime. In the case
of PFOA, the POD is the human equivalent dose (HED) derived from the modeled serum
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concentration representing either an NOAEL or LOAEL experimental dose after applying
uncertainty factors established following EPA guidelines.
RfD = HEP NOAEL or HEP LOAEL
UF
Where:
HED NOAEL = The HED from the modeled average serum representing the highest of the
given doses that lacked adverse effects (mg/kg/day).
HED LOAEL = The HED from the modeled average serum representing the lowest of the
given doses that results in adverse effects (mg/kg/day) and of an appropriate duration
and endpoint to use for a lifetime HA.
UF = Total Uncertainty Factor established in accordance with EPA guidelines
considering variations in sensitivity among humans, differences between animals and
humans, the duration of exposure in the key study compared to a lifetime of the
species studied, whether the HED is a dose that caused an effect or no effect, and the
completeness of the toxicology database.
Step 2: Calculate a Drinking Water Equivalent Level (DWEL) from the RfD. The DWEL
assumes that 100% of the exposure comes from drinking water.
RfD x bw
Where:
RfD = Reference dose (mg/kg bw/day)
bw = Assumed body weight (kg)
DWI = Assumed human daily drinking water intake (L/day)
Step 3: Calculation of the Lifetime HA. The lifetime HA is calculated by factoring in other
sources of exposure (e.g., air, food, soil) in addition to drinking water using the methodology
described for calculation of an RSC described in USEPA (2000) and section 6.1.
Lifetime HA = DWEL x RSC
Where:
DWEL = Drinking water equivalent level calculated from step 2 (mg/L)
RSC = Relative source contribution
3.2.4 Measures of Effect
The animal toxicology studies were used in the dose-response assessment for PFOA. These
studies demonstrated dose-related effects on systemic and developmental endpoints in multiple
species (monkeys, rats, mice) following exposure to PFOA for durations of 1 1 to 84 days; these
are described in detail in the HESD for PFOA. The studies selected for pharmacokinetic analysis
were chosen based on their experimental design, data quality, dose-response data identified
through the range of experimental NOAELs/LOAELs, and serum measurements of PFOA.
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EPA used a peer-reviewed pharmacokinetic model developed by Wambaugh et al. (2013) to
calculate the average serum concentrations associated with the candidate NOAELs and LOAELs
from the toxicological database. Average serum levels of PFOA from the model were used to
determine the HED associated with the study NOAEL and LOAEL. The Wambaugh et al. (2013)
model is based on the Andersen et al. (2006) concept that saturable renal resorption is
responsible for the long serum half-lives seen in humans and animals.
A unique feature of the pharmacokinetic approach is the use of a single model for the three
species and reliance on the serum PFOA level as the measure of exposure. For each species the
model accommodated the appropriate toxicokinetic variables for the species/strain. The
pharmacokinetic analysis facilitated examination for consistency in the average serum values
associated with effect and no-effect doses from the animal PFOA studies. A nonhierarchical
model for parameter values was assumed wherein a single numeric value represented all
individuals of the same species, gender, and strain. Body weight, the number of doses, and
magnitude of the doses were the only parameters that varied.
3.2.5 Relative Source Contribution
The RSC is applied in the HA calculation to ensure that an individual's total exposure from a
contaminant (i.e., PFOA) does not exceed the RfD. The RSC is the portion of the RfD attributed
to drinking water (directly or indirectly in beverages like coffee, tea, or soup); the remainder of
the RfD is allocated to other potential exposure sources. In the case of PFOA, other potential
sources include ambient air, foods, incidental soil/dust ingestion, consumer products, and others
(see sections 2.2 and 6.1). The RSC for the HA is based on exposure to the general population.
EPA derived an RSC for PFOA by using the Exposure Decision Tree approach (USEPA
2000) (see section 6.1). To use that approach, EPA compiled information for PFOS on its uses,
chemical and physical properties, occurrences in other potential sources (e.g., air, food), and
releases to the environment. To determine the RSC to be used in the HA calculation for PFOA,
EPA then used the information to address the questions posed in the Exposure Decision Tree.
Some of the important items evaluated in the Exposure Decision Tree are:
• Adequacy of data available for each relevant exposure source and pathway.
• Availability of information sufficient to characterize the likelihood of exposure to
relevant sources.
• Whether there are significant known or potential uses/sources other than the source of
concern (i.e., ambient water and fish/seafood from those waters).
• Whether information on each source is available to characterize exposure.
In cases where environmental or exposure data are lacking, the Exposure Decision Tree
approach results in a recommended RSC of 20%. This 20% RSC value may be replaced where
sufficient data are available to develop a scientifically defensible alternative value. When
appropriate, if scientific data demonstrating that sources and routes of exposure other than
drinking water are not anticipated for the pollutant in question, the RSC may be raised to 80%
based on the available data (USEPA 2000).
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4.0 EFFECTS ASSESSMENT
The database for PFOA includes a large number of laboratory animal toxicity studies, as well
as numerous epidemiology studies. The most extensive epidemiology studies were conducted by
the C8 Science Panel for a highly exposed population in West Virginia. These animal and human
studies are described below and in greater detail in the HESD for PFOA (USEPA 2016a).
Because of uncertainties associated with the human data (described above), EPA is relying on
animal data to quantitatively assess effects; however, the epidemiology studies provide important
data to establish probable links between PFOA exposure to humans and health effects. In
particular, effects on the liver enzymes indicative of liver effects, low birthweight, antibody
response, and cancer in laboratory animals are supported by human epidemiology studies.
4.1 Noncancer Health Effects
4.1.1 Animal Toxicity Studies
The database of animal toxicology studies is extensive with short term, subchronic, and
chronic toxicity and cancer studies; developmental and reproductive toxicity, neurotoxicity, and
immunotoxicity studies; and mechanistic studies.
Developmental Effects
Both rats and mice showed developmental toxicity based on low birth weights, skeletal
effects (reduced ossification), altered onset of puberty (Butenhoff et al. 2004a; Lau et al. 2006;
Wolf et al. 2007). Doses that elicited a response were higher in rats compared with those in mice.
Meta-analyses were conducted to determine whether developmental exposure to PFOA was
associated with fetal growth effects in animals (Koustas et al. 2014). Eight animal studies
identified in the published literature met the criteria of the Navigation Guide systematic review
methodology as developed and published by Woodruff and Sutton (2014) for inclusion in the
analyses. The animal data sets included mouse gavage studies with maternal PFOA doses from
0.01 to 20 mg/kg/day. The results from the meta-analysis showed that a 1 mg/kg/day increase in
PFOA dose was associated with a -0.023 g (95% CI [-0.029, -0.016]) difference in pup birth
weight. The MOA for decreased pup body weight is not known, but receptor-activated changes
in metabolism, hormonal perturbations, and impeded intercellular communication might play a
role.
One animal neurological study (Johansson et al. 2009) showed effects on habituation and
activity patterns in NMRI (Naval Medical Research Institute) mice treated on post-natal day
(PND)IO with a single dose of PFOA and evaluated at 2 and 4 months of age (LOAEL =
0.58 mg/kg). The in vivo observations are supported by changes in the expression of a variety of
neurologically active brain proteins in the treated pups (Johansson et al. 2009). The offspring of
C57BL/6/Bkl dams fed diets that provided a dose of 0.3 mg PFOA/kg/day throughout gestation
had detectable levels of PFOA in their brains at birth (Onishchenko et al. 2011). Behavioral
assessments of the offspring starting at 5 weeks of age revealed sex-related differences in
exploratory behavior patterns. In the social group setting, the PFOA-exposed males were more
active and PFOA-exposed females were less active than their respective controls. The PFOA-
exposed males also had increased activity counts compared to control males in circadian activity
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experiments. The results of an in vitro study of hippocampal synaptic transmission and neurite
growth in the presence of long chain perfluorinated compounds showed that 50 or 100
micromolar PFOA increased spontaneous synaptic current and had an equivocal impact on
neurite growth (Liao et al. 2009a, 2009b). These data suggest a need for additional studies of the
effects of PFASs, including PFOA, on the brain.
The developmental impacts of PFOA exposure ranged from delayed mammary gland
development in pups (Albrecht et al. 2013; Macon et al. 2011; Tucker et al. 2015; White et al.
2009, 2011; Wolf et al. 2007) to delays in attaining developmental milestones (Lau et al. 2006;
White et al. 2009; Wolf et al. 2007). The LOAEL for the mammary gland developmental effects
in female offspring from dams given 0.01 mg/kg/day for 8 days from Macon et al. (2011) is of
unknown biological significance. The same study showed no effects on offspring body weight at
maternal doses up to 3 mg/kg/day for 17 days (Macon et al. 2011). Data from White et al. (2011)
showed no significant effects on body weight gain in pups nursing from dams treated with 1
mg/kg/day, despite these dams having less fully developed mammary glands compared to
controls. Similarly, no differences in response to a lactational challenge were seen in PFOA
exposed dams with morphologically delayed mammary gland development (White et al. 2011).
Immune Function
Several animal studies demonstrate effects on the spleen and thymus as well as their cellular
products (B lymphocytes and T-helper cells) in several strains of mice. Studies by Yang et al.
(2000, 2001, 2002b) and DeWitt et al. (2008) were conducted using relatively high PFOA doses
(-30 to 40 mg/kg/day). In each study, the PFOA-treated animals exhibited significant decreases
in spleen and thymus weights as well as splenocyte and thymocyte populations at various stages
of differentiation. Recovery usually occurred within several days of cessation of PFOA dosing.
When the response of C57BL/6Tac PPARa mice were compared to wild type of the same strain,
the knockout mice showed no response with both spleen and thymus weights at 30 mg/kg/day,
whereas there was a response in the wild-type strain (DeWitt et al. 2015), suggesting an impact
of PPARa. Both strains showed an increase in immunoglobulin M (IgM) in response to a sheep
red blood cell (SRBC) injection. The 30 mg/kg/day dose was the LOAEL for the knockout mice
and 7.5 mg/kg/day was the response level for the wild-type strain. Thus, the suppression of the
immune system is not totally a PPARa-related response.
DeWitt et al. (2008) used different functionality assays in their study in C57B1/6 mice. The
IgM response to SRBCs was suppressed by 20% when mice were immunized immediately after
exposure to the initial dose of 30 mg PFOA/kg/day ceased. However, no significant increase
occurred in the response to bovine serum albumin 4 days post-PFOA exposure, or in the
immunoglobulin G (IgG) response to SRBC 15 days post-PFOA exposure. These results are
indicative of recovery once PFOA exposures have ceased. DeWitt et al. (2008) followed their
initial study of PFOA with one designed to examine the dose response for a 15-day drinking
water exposure in a slightly different mouse strain, C57B1/6N. The LOAEL was 3.75 mg/kg/day
based on a significant decrease in IgM response, and the NOAEL was 1.88 mg/kg/day indicating
the inability to respond to an immunological challenge.
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Liver Disease and Liver Function
Hepatocellular hypertrophy and an increased liver-to-body weight ratio are common findings
in rodents, but are considered non-adverse if there is evidence for PPARa activation. These
effects are considered adverse if accompanied by necrosis, fibrosis, inflammation, and steatosis
(Hall et al. 2012). Low-level necrotic cell damage was observed in the Palazzolo et al. (1993) rat
study and in the Loveless et al. (2008) studies in CD rats and CD-I mice. Palazzo et al. (1993) is
an unpublished report that was later published as Perkins et al. (2004). The liver histopathology
details of this study were only presented in Palazzolo et al. (1993). This study will be referred to
throughout the rest of the document as Palazzolo et al. (1993)/Perkins et al. (2004). In this study
there was a slight increase in coagulative necrosis at 1.94 and 6.5 mg/kg/day when compared to
the control and lower dose (0.94 mg/kg/day). Some hepatocellular necrosis was also observed in
conjunction with hepatocellular hypertrophy and increased liver weight at a dose of 3 mg/kg/day
in Fl male rats from the Butenhoff et al. (2004a) two-generation study.
In general, effects on organs other than the liver tend to occur at doses higher than those that
affect the liver. Lung effects including pulmonary congestion were observed in male Sprague-
Dawley rats (LOAEL = 5 mg/kg/day) by Cui et al. (2009). Increased thickness and prominence
of the adrenal zona glomerulosa and vacuolization in the cells of the adrenal cortex were
observed in male rats fed 10 mg/kg/day for approximately 56 days (Butenhoff et al. 2004a).
Kidney Function
Some studies have shown effects on the kidney of male rats at doses similar to those resulting
in liver effects. Increases in absolute and relative-to-body kidney weights occurred in rats given
5 mg/kg/day (lowest dose tested) via gavage for 28 days (Cui et al. 2009). In a two-generation
gavage study, FO and Fl males had significantly increased absolute kidney weight at 1 and
3 mg/kg/day, but significantly decreased kidney weight at 30 mg/kg/day. Organ weight-to-
terminal body weight ratios for the kidney were statistically significantly increased at > 1
mg/kg/day. Kidney weight-to-brain weight ratios were significantly increased at 1, 3, and
10 mg/kg/day, but decreased at 30 mg/kg/day (Butenhoff et al. 2004a). In the high-dose group,
absolute and relative kidney weight changes occurred in a pattern typically associated with
decrements in body weight and are indicative of systemic toxicity. In the lower-dose groups, the
consistently increased absolute and relative to body and brain weights suggest a cellular
response, whereby the kidney tubular cells upregulate expression of transporter proteins to
facilitate the PFOA excretion. This is adverse because it is a biomarker for systemic PFOA
bioaccumulation. The differential expression of transporters in the kidney of male rats is under
hormonal control with males having lower levels of export transporters compared to females
(Kudo et al. 2002). No dose-related changes in kidney weight or histopathology were found in
male rats at the end of 2 years with a dose of 14.2 mg/kg/day (Butenhoff et al. 2012).
Diabetes
Hines et al. (2009) found no differences in glucose tolerance tests at 15-16 weeks and at 17
months of age in PFOA-exposed CD-I mice, but did observe significantly increased serum leptin
and insulin levels at 21 and 31 weeks of age suggesting that the insulin resistance mechanistic
pathway could be affected by PFOA and a connection between PFOA and increased body
weight. Leptin is a hormone secreted by adipose tissue that is associated with weight gain.
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Conversely, Quist et al. (2015) found no dose-related impact on serum leptin in CD-I pups on
PND 91. Quist et al. (2015) found that when mice were on a high-fat diet and not fasted before
serum collection and these were compared to the same mice that were fasted before serum
collection, leptin increased, thereby suggesting that the leptin change could be temporary and
dependent on the fat content of the diet and the timing of serum collection.
Thyroid
Effects of PFOA on thyroid hormones in animals are not as well characterized as those of
PFOS. Butenhoff et al. (2002) evaluated the toxicity of PFOA in a small number of male
monkeys during 6 months of oral administration and reported that levels of total T3 and free T3
in circulation were reduced significantly in the 30/20 mg/kg/day treatment group, beginning at
5 weeks after initiation of treatment but accompanied by other signs of systemic toxicity.
Recovery of T3 deficits was noted when PFOA returned to baseline 90 days later. Serum total
thyroxine (T4), free T4, and thyroid-stimulating hormone (TSH) were not altered throughout the
study. The preferential effects of PFOA on serum T3 and a lack of a TSH compensatory
response are similar to those observed with PFOS, and are possibly a consequence of PFOA
binding to the T3 receptor (Ren et al. 2015). None of the thyroid hormones were affected by
PFOA in mature female rats (Butenhoff et al (2002), primarily because these animals were able
to clear the chemical effectively (with half-life estimate of 2 to 4 hours, compared to that of 6 to
7 days for male rats). This suggests that the thyroid disrupting effects of PFOA are directly
related to endogenous accumulation of the chemical and might be relevant to humans because of
the long PFOA human half-life.
Fertility, Pregnancy, and Birth Outcomes
Among animal studies there was no effect of PFOA on reproductive or fertility parameters in
rats (Butenhoff et al. 2004a), but effects on male fertility were observed in mice given a dose of
5 mg/kg/day for 28 days prior to mating (Lu et al. 2015). A NOAEL of 2.5 mg/kg/day and a
LOAEL of 5 mg/kg/day were reported for reduced sperm counts and changes in testicular
morphology after a 14-day exposure by Liu et al. (2015); 2 mg/kg/day led to significantly
increased serum estradiol and increased hepatic aromatase activity in the same study. Gender
differences in dose response are likely related to half-life differences of hours for the female rat
and days-to-weeks for the female mouse.
Serum Lipids
Information on serum lipids from animal studies has received less attention than in the
human population because decreases in triglycerides, cholesterol, and lipoprotein complexes are
an expected consequence of PPARa activation in rodents. PFOA is an activator of the PPARa
nuclear receptor in both humans and animals, but activation in humans does not increase the
cellular levels of peroxisomes to the same extent it does in rodents. The PPARa response in
animals tends to lower rather than raise serum cholesterol and associated lipid levels. PFOA is
known to activate the PPAR pathway by increasing transcription of mitochondrial and
peroxisomal lipid metabolism, sterol, and bile acid biosynthesis and retinol metabolism genes.
However, based on transcriptional activation of many genes in PPARa null mice, the effects of
PFOA involve more than activation of PPAR. Also activated are the constitutive androstane
receptor (CAR), farnesoid X receptor (FXR), and pregnane X receptor (PXR).
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Cholesterol and/or triglycerides were monitored in few animal studies. Nakamura et al.
(2009) found that mice with a normal PPARa receptor had significantly increased levels of
cholesterol and triglycerides in liver, but not plasma, at a LOAEL of 0.3 mg/kg/day. However,
no differences were observed in serum or liver cholesterol or triglycerides between PFOA-
treated mice with a humanized PPARa receptor or PPARa null mice (NOAEL = 0.3 mg/kg/day)
and their respective controls. A study by Minata et al. (2010) used higher doses and found that
total cholesterol was significantly decreased and total triglycerides significantly increased in
wild-type mice. In the PPARa null mice, total triglycerides were significantly increased at all
doses.
In animal studies, serum levels of alanine aminotransferase (ALT) and/or aspartate
aminotransferase (AST) were significantly increased indicative of apoptosis or necrosis of liver
cells (Butenhoff et al. 2012; Minata et al. 2010; Son et al. 2008). Increased levels of ALT were
observed at a LOAEL of 2.65 mg/kg/day in ICR mice by Son et al. (2008). Yahia et al. (2010)
reported significantly increased ALT, gamma-glutamyl transferase (GGT), AST, and alkaline
phosphatase (ALP) in PFOA-exposed (10 mg/kg) pregnant ICR mice. Total protein, albumin,
and globulin were significantly decreased in the same mice.
4.1.2 Human Epidemiology Studies
Numerous epidemiology studies evaluating large cohorts of highly exposed occupational and
general populations have examined the association of PFOA exposure to a variety of health
endpoints. Health outcomes assessed include blood lipid and clinical chemistry profiles;
reproductive parameters; thyroid effects; diabetes; immune function; birth, fetal, and
developmental growth measures; and cancer. Members of the general population living in the
vicinity of the West Virginia DuPont Washington Works PFOA production plant in Parkersburg,
West Virginia, are the focus of an ongoing study titled the C8 Health Project. Releases from the
Washington Works plant, where PFOA was used as a processing aid in the manufacture of
fluoropolymers, contaminated the ground water from six water districts near the plant, resulting
in exposures to the general population. The C8 Health Project is the largest study evaluating
human exposure and health endpoints for PFOA; the study included more than 65,000 people in
Mid-Ohio Valley communities who were exposed to PFOA for longer than 1 year.
As part of the C8 Health Project, a panel of expert epidemiologists reviewed the
epidemiological and other data available in 2011 and 2012 to assess probable links between
PFOA exposure and disease.6 The C8 Science Panel concluded that a probable link existed
between PFOA exposure and the following conditions: high cholesterol, thyroid disease,
pregnancy-induced hypertension, ulcerative colitis, and kidney and testicular cancer. The C8
Science Panel did not find a probable link between PFOA exposure and multiple other
conditions, including other autoimmune diseases (rheumatoid arthritis, lupus, Type I diabetes,
Crohn's disease, multiple sclerosis), Type II diabetes, high blood pressure, coronary artery
disease, infectious disease, liver disease, Parkinson's disease, osteoarthritis, neurodevelopmental
disorders in children (attention deficit hyperactivity disorder, learning disabilities), chronic
kidney disease, stroke, asthma or chronic obstructive airways disease (COPD), and birth defects,
6 For more information see http://www.c8sciencepanel.org/prob link.html.
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miscarriage or stillbirth, preterm birth or low birth weight, and other types of cancer. The
summary below focuses on the endpoints highlighted as "probable links" by the C8 Science
panel, and on other epidemiology studies published after the 2011-2012 reports.
Serum Lipids
The association between PFOA and serum lipids has been examined in several studies in
different populations. Cross-sectional and longitudinal studies in occupational settings (Costa et
al. 2009; Olsen et al. 2000, 2003; Olsen and Zobel, 2007; Sakr et al. 2007a, 2007b; Steenland et
al. 2015) and in the high-exposure community (the C8 Health Project study population) (Fitz-
Simon et al. 2013; Frisbee et al. 2010; Steenland et al. 2009; Winquist and Steenland 2014a)
generally observed positive associations between serum PFOA and total cholesterol in adults and
children (ages 1 to < 18 years); most of these effect estimates were statistically significant.
Although exceptions to this pattern are present (i.e., some of the analyses examining incidence of
self-reported high cholesterol based on medication use in Winquist and Steenland, 2014a, and in
Steenland et al. 2015), the results are relatively consistent and robust. Similar associations were
seen in analyses of low-density lipoprotein (LDL), but were not seen with high-density
lipoprotein (HDL). The range of exposure in occupational studies is large (means varying
between 0.4 and > 12 micrograms per milliliter [|ig/mL]), and the mean serum levels in the C8
population studies were around 0.08 jig/mL. Positive associations between serum PFOA and
total cholesterol (i.e., increasing lipid level with increasing PFOA) were observed in most of the
general population studies at mean exposure levels of 0.002 to 0.007 |ig/mL (Eriksen et al. 2013;
Fisher et al. 2013; Geiger et al. 2014; Nelson et al. 2010; Starling et al. 2014). The interpretation
of these general population results is limited, however, by the moderately strong correlations
(Spearman r > 0.6) and similarity in results seen for PFOS and PFOA.
Liver Disease and Liver Function
Few studies pertaining to the relation between PFOA and liver disease are available. The C8
Health Project did not observe associations with hepatitis, fatty liver disease, or other types of
liver disease. In the studies of PFOA exposure and liver enzymes (measure in serum), positive
associations were seen. The results of the occupational studies provide evidence of an association
with increases in serum AST, ALT, and GGT, with the most consistent results seen for ALT. The
associations were not large, and the associations could depend on the co-variates in the models,
such as body mass index, use of lipid lowering medications, and triglycerides (Costa et al. 2009;
Olsen et al 2000, 2003; Olsen and Zobel, 2007; Sakr et al. 2007a, 2007b).
Two population-based studies of highly exposed C8 area residents evaluated associations
with liver enzymes, and the larger of the two studies reported associations of increasing serum In
ALT and In GGT levels with increasing serum PFOA concentrations (Emmett et al. 2006; Gallo
et al. 2012). A cross-sectional analysis of data from NHANES, representative of the U.S.
national population, also found associations with In PFOA concentration with increasing serum
ALT and In GGT levels. Serum bilirubin was inversely associated with serum PFOA in the
occupational studies. A U-shaped exposure-response pattern for serum bilirubin was observed
among the participants in the C8 Health Project which might explain the inverse associations
reported for occupational cohorts. Overall, an association of serum PFOA concentration with
elevations in serum levels of ALT and GGT was consistently observed in occupational, highly
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exposed residential communities, and the U.S. general population. The associations are not large
in magnitude, but indicate the potential to affect liver cells.
Immune Function
Three studies examined associations between maternal and/or child serum PFOA levels and
vaccine response (measured by antibody levels) in children (Grandjean et al. 2012; Granum et al.
2013) and adults (Looker et al. 2014). The study in adults was part of the high-exposure
community C8 Health Project; a reduced antibody response to one of the three influenza strains
tested after receiving the flu vaccine was seen with increasing levels of serum PFOA. The studies
in children were conducted in general populations in Norway and in the Faroe Islands. As
observed in the animal studies, decreased vaccine response in relation to PFOA levels was seen
in these studies, but similar results also were seen with other correlated PFASs (e.g., PFOS).
Thyroid
Three studies reported an increased risk of thyroid disease in women or girls, but not men
(Lopez-Espinosa et al. 2012; Melzer et al. 2010; Winquist and Steenland, 2014b). A fourth study
also reported a trend of elevated TSH and decreased T4 (hypothyroidism) in pregnant women
testing positive for hypothyroid autoimmune disease (Webster et al. 2014). Similarly, the C8
Panel concluded there was strong evidence to link PFOA exposure to thyroid disease in its
population. Hypothyroxinemia (decreased free thyroxine (FT4) without concomitant elevation of
TSH) was measured in one study of pregnant women showing null findings for
hypothyroxinemia incidence versus controls; hypothyroxinemia is not typically studied in the
clinic as TSH and T4 concomitantly inversely shift with thyroid disease. Looking at thyroid
hormone levels, some studies found changes in thyroid hormone levels associated with PFOA
(de Cock et al. 2014; Shrestha et al. 2015; Webster et al. 2014; Wen et al. 2013,); others found
null effects of PFOA in association with thyroid hormones. Generally null associations were
found in other studies on the general population, pregnant women, and patients in association
with thyroid hormone levels or one portion of the thyroid panel was outside of control range.
Across multiple studies, thyroid hormone concentrations have mixed evidence (associations and
null findings) in association with PFOA concentrations. Increased risk for thyroid disease in
women appears to be associated with PFOA serum concentration; evidence is weaker or null in
men.
Diabetes
No associations were observed between serum PFOA levels and type II diabetes incidence
rate in general or worker populations with mean serum PFOA up to 0.0913-0.113 |ig/mL
(MacNeil et al. 2009; Steenland et al. 2015). PFOA was not associated with measures of
metabolic syndrome in adolescents or adults (Lin et al. 2009). However, one study found an
increased risk for developing gestational diabetes in women with mean serum PFOA (measured
at preconception) of 0.00394 |ig/mL (Zhang et al. 2015).
Fertility, Pregnancy, and Birth Outcomes
The association between PFOA and birth weight has been examined in numerous studies (see
section 4.1.1.7 in USEPA 2016a). Most studies measured PFOA in the general population using
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maternal blood samples taken in the second or third trimester or in cord blood samples. One
study was able to collect samples earlier in the pregnancy (4-14 weeks) (Fei et al. 2007), and
another study in the high-exposure community (the C8 Health Project population) modeled
exposure based on data on residential history and environmental data (Savitz et al. 2012). Two
meta-analyses of these studies have been conducted (Johnson et al. 2014; Verner et al. 2015),
with similar results: mean birthweight reduction of 19 g (95% CI [-30, -9]) per each one unit
(ng/mL) increase in maternal or cord serum PFOA levels in Johnson et al. (2014), and a mean
birthweight reduction of 15 g (95% CI [-22, -8]) based on seven of these nine studies in Verner
et al. (2015). It has been suggested that low glomerular filtration rate (GFR) can affect birth
weight (Morken et al 2014). Verner et al (2015) conducted a meta-analysis based on
physiologically based pharmacokinetic model (PBPK) simulations and found that some of the
association reported between PFOA and birth weight is attributable to GFR and that the actual
association may be closer to a 7 gram reduction (95% CI [-8, -6]). Verner et al. (2015) showed
that in individuals with low GFR there are increased levels of serum PFOA and lower birth
weights. Although some uncertainty exists in the interpretation of the observed association
between PFOA and birth weight given the potential impact of low GFR, the available
information indicate that the association between PFOA exposure and birth weight cannot be
ruled out. In humans with low GFR (which includes women with pregnancy-induced
hypertension or preeclampsia) the impact on body weight is likely due to a combination of the
low GFR and the serum PFOA.
Two studies examined development of puberty in girls in relation to prenatal exposure to
PFOA as measured through maternal or cord blood samples in follow-up of pregnancy cohorts
conducted in England (Christensen et al. 2011) and in Denmark (Kristensen et al. 2013). The
results of these two studies are conflicting, with no association (or a possible indication of an
earlier menarche seen with higher PFOA) in Christensen et al. (2011), and a later menarche seen
with higher PFOA in Kristensen et al. (2013). Another study examined PFOA exposure
measured concurrently with the assessment of pubertal status (Lopez-Espinosa et al. 2011). An
association between later age at menarche and higher PFOA levels was observed, but the
interpretation of this finding is complicated by the potential effect of puberty on the exposure
biomarker levels (i.e., reverse causality). Menstruation is a route of excretion for albumin-bound
PFOA; thus, the beginning of menstruation will remove serum PFOA when the menstruation
periods begin during puberty and its cessation at menarche will decrease the loss of PFOA in
blood and allow serum levels to increase.
Limited data suggest a correlation between higher PFOA levels (>0.02 |ig/mL) in women
and decreases in fecundity and fertility (Fei et al. 2009; Velez et al. 2015), but there are no clear
effects of PFOA on male fertility endpoints (0.0035-0.005 |ig/mL; Joensen et al. 2009, 2013).
4.1.3 Noncancer Mode of Action
No published cohesive MOA exists that accounts for the varied toxicological properties of
PFOA. However, a number of the unique properties of the compound contribute to its toxicity:
• Metabolic stability accompanied by persistence in tissues as an apparent consequence of
saturable renal resorption.
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• Electrostatic binding to biopolymers with areas of positive charge, especially proteins
(MacManus-Spencer et al. 2009; Salvalaglio et al. 2010; Wu et al. 2009b; L. Zhang et al.
2013).
• Displacement of endogenous/exogenous substances normally bound to serum albumin
such as fatty acids, bile acids, pharmaceuticals, and T3 (Fasano et al. 2005; Qin et al.
2010; Wuetal. 2009a).
• Renal resorption (Andersen et al. 2006) and biliary excretion that are dependent on
transporters genetically encoded for management of natural substances (endogenous and
exogenous) that prolong systemic retention of absorbed PFOS and explain its long half-
life
• Binding to and activating receptors such as PPAR, thereby initiating activation or
suppression of gene transcription related to fatty acid metabolism and lipid transport
(Nakamura et al. 2009; Rosen et al. 2007, 2009a, 2009b; Takacs and Abbott 2007).
• Interference with intercellular communication (Upham et al. 1998, 2009).
The renal resorption and biliary competition between natural substrates and PFAS contribute
to ambiguity in some of the epidemiology study outcomes where serum levels of endogenous or
dietary-transported substrates are altered because of preferential removal or resorption of the
PFOA, or the PFOA serum level increases because of the preferred excretion of the natural
material. Physiological status also has an impact on the epidemiology results given that blood
loss through menstruation is an excretory pathway for serum-albumin-bound PFOA. Thus, serum
levels will be lower in girls after puberty than before, and will increase in women after
menopause. In pregnant women, increased blood volume as well as cessation of monthly
menstrual blood flow route also influences serum levels.
The outcome from studies of antibodies or immunoglobulins can be confounded by PFOA
protein binding depending on the impact of morphological changes caused by PFOA binding on
the sensitivity of the assay. Interaction of PFOA and other PFASs (Ren et al. 2015) with the T3
receptor has the potential to influence cellular uptake of T3. Binding to thyroid hormone
transport protein or transthyretin (TTR) can displace T4 increasing the unbound level in serum
(Weiss et al. 2009).
There are no cohesive studies designed to identify modes of action for the liver weight and
hypertrophy endpoints represented in the animal studies. Both effects are clearly, in part, an
outcome of PPARa activation. They become adverse when accompanied by inflammation,
fibrosis, steatosis, or necrosis (Hall et al. 2012) as seen in Palazzolo et al. (1993)/Perkins et al.
(2004), Loveless et al. (2008), and Butenhoff et al. (2004a).
The MOA for decreased pup body weight observed in the animal studies is unknown
(Butenhoff et al. 2004a; White et al. 2009; Wolf et al. 2007). The observed effects on birth
weight in animals are supported by evidence of an association between PFOA and low birth
weight in humans (Johnson et al. 2014). Receptor-activated changes in metabolism, hormonal
perturbations, and impeded intercellular communication could play a role in this effect. It has
been suggested that GFR can affect birth weight (Morken et al 2014). Verner et al (2015)
conducted a meta-analysis based on PBPK simulations and found that some of the association
reported between PFOA and birth weight could be partially attributable to low GFR and that the
actual association might be closer to a 7 gram reduction (95% CI [-8, -6]). However, the study
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) - May 2016 43
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authors demonstrated that individuals with low GFR have increased levels of serum PFOA and
lower birth weights. Although some uncertainty exists in the interpretation of the observed
association between PFOA and birth weight given the potential impact of low GFR, the available
information indicate that the association between PFOA exposure and birth weight cannot be
ruled out. In humans with low GFR (which includes women with pregnancy-induced
hypertension or preeclampsia) the impact on body weight is likely due to a combination of the
low GFR and the serum PFOA.
Women with hypertension during pregnancy are a susceptible population that could have an
increased risk for having a low birth weight baby.
There also is a lack of data relative to the MOA for immunological effects of PFOA as seen
in animal studies. Some of the responses are PPARa linked (increased spleen and thymus
weights) but not all as demonstrated by DeWitt et al. (2015). Effects on serum immunoglobulins
observed in humans could be a reflection of analytical method interference as a result of PFOA
binding to the immunoglobin (Kerstner-Wood et al. 2003).
PFOA is associated with delayed breast tissue development (reduced ductal branching and
numbers of terminal endbuds) in CD-I mice (Albrecht et al. 2013; Macon et al. 2011; Tucker et
al. 2015; White et al. 2009); however, no functional impacts on the ability of the dams to provide
nourishment were observed based on the weight increases in the pups reared by the impacted
dams (Macon et al. 2011; White et al. 2011). CD-I mice seem to be more sensitive for this effect
than other mice strains evaluated (Tucker et al. 2015). No mechanistic studies exist that inform
the MOA for the mammary gland development effects.
Quist et al. (2015) found that the level of dietary fat in an animal diet is an important variable
that influences liver lipid levels. At PFOA doses < 0.3 mg/kg/day, the LDL and total serum
cholesterol levels in the fasted and nonfasted high-fat diet animals were greater than in the
untreated Purina controls. Tan et al. (2013) found that the fat content of the diet was an important
variable in determining the impact of PFOA (5 mg/kg/day) on liver and serum lipids. Intake of a
high-fat diet plus PFOA increased liver triglycerides and serum free fatty acids as compared to a
regular fat diet plus PFOA, but it had no impact on liver cholesterol concentrations. Serum
cholesterol was not monitored. A high-fat diet predisposes animals and possibly humans to
hepatic steatosis.
4.2 Cancer
4.2.1 Animal Cancer Bioassays
The only animal carcinogenicity studies available for PFOA indicate that exposure can lead
to liver adenomas (Biegel et al. 2001), Ley dig cell adenomas (Biegel et al. 2001; Butenhoff et al.
2012), and pancreatic acinar cell tumors (PACT) (Biegel et al. 2001) in male Sprague-Dawley
rats. In the Butenhoff et al. (2012) study there was an increase in liver carcinomas at the high
dose (14.2 mg/kg/day) in the males compared to controls (6% versus 10%). For the females
receiving 16.1 mg/kg/day (i.e., the high dose) the tumor incidence compared to controls was 0%
versus 2%. The increase in liver tumors did not show a direct relationship to dose in the male rats
and was not statistically significantly elevated in either males or females at the high dose when
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compared to controls (Butenhoff et al. 2012). Liver adenomas were observed in the Biegel et al.
(2001) study at an incidence of 10/76 (13%) at 20 mg/kg/day. The incidence in the control group
was 2/80 (3%).
Butenhoff et al. (2012) also observed increased incidence of testicular Leydig cell tumors
(LCTs) in rats. At the 1-year sacrifice, testicular masses were found in 7/50 (14%) high-dose and
2/50 (4%) low-dose rats, but not in any of the controls. A significant increase (p < 0.05) in the
incidence of testicular (Leydig) cell adenomas was also observed in the high-dose male rats at
the end of the study. The LCT incidence in the control, low-, and high-dose groups was 0/50
(0%), 2/50 (4%), and 7/50 (14%), respectively. Biegel et al. (2001) observed a significant
increase in the incidence of Leydig cell adenomas in the treated rats (11%, 8/76) when compared
to the pair-fed control rats (3%, 2/78) supporting the observations from the Butenhoff et al.
(2012) study. The LCTs in the Butenhoff et al. (2012) study were accompanied by statistically
significant testicular vascular mineralization and by Leydig cell hyperplasia in the Biegel et al.
(2001) study.
PACTs were only observed in the Biegel et al. (2001) study, with an incidence of 11% at
20 mg/kg/day compared to controls. Although no PACTs were observed by Butenhoff et al.
(2012), pancreatic acinar hyperplasia was observed at 1.3 and 14.2 mg/kg/day at incidences of
6% and 2%, respectively. Re-examination of the pancreatic lesions in Butenhoff et al. (2012) and
Biegel et al. (2001) resulted in the conclusion that the high dose increased the incidence of
proliferative acinar cell lesions in both studies. Some lesions in the Biegel et al. (2001) study had
progressed to adenomas but not those in the Butenhoff et al. (2012) study.
The initial findings from the Butenhoff et al. (2012) study were equivocal for mammary
fibroadenomas in female rats. However, a re-examination of the tissues by a pathology working
group (PWG) found no statistically significant differences in the incidence of fibroadenomas or
other neoplasms of the mammary gland between control and treated animals (Hardisty et al.
2010). The PWG used the diagnostic criteria and nomenclature of the Society of Toxicological
Pathologists for the re-examination.
4.2.2 Human Epidemiology Studies
Evidence of carcinogenic effects of PFOA in epidemiology studies is based on studies of
kidney and testicular cancer. These cancers have relatively high survival rates (2005-2011
5-year survival rates are 73% and 95%, respectively, for kidney and testicular cancer, based on
NCI Surveillance, Epidemiology and End Results data); therefore, studies that examine
population cancer incidence are particularly useful for these types of cancers. For testicular
cancer, the high-exposure community studies also have the advantage of including the age period
of greatest risk, as the median age at diagnosis is 33 years. The two occupational cohorts in
Minnesota and West Virginia (most recently updated, respectively, in Raleigh et al. 2014 and
Steenland and Woskie, 2012) do not support an increased risk of these cancers, but each of these
is limited by a small number of observed cases (six kidney cancer deaths, 16 incident kidney
cancer cases, and five incidence testicular cancer cases in Raleigh et al. [2014]; 12 kidney cancer
deaths and one testicular cancer death in Steenland and Woskie [2012]). Two studies involving
members of the C8 Health Project showed a positive association between PFOA levels (mean at
enrolment 0.024 |ig/mL) and kidney and testicular cancers (Barry et al. 2013; Vieira et al. 2013);
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) - May 2016 45
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some of the cases included in these studies overlap. None of the general population studies
examined kidney or testicular cancer, but no associations were found in the general population
between mean serum PFOA levels up to 0.0866 |ig/mL and colorectal, breast, prostate, bladder,
and liver cancer (Bonefeld-J0rgensen et al. 2014; Eriksen et al. 2009; Hardell et al. 2014; Innes
etal. 2014).
4.2.3 Cancer Mode of Action
The mode of carcinogenic action of PFOA is not clearly understood. Some researchers have
concluded from the available data that the liver tumors observed in the cancer bioassays can be
attributed mostly to the impact of PFOA on peroxisome proliferation based on a hypothesized
lower sensitivity of humans to this MOA (Klaunig et al. 2003, 2012). Some data support the
hypothesis that PPARa agonism MOA could be responsible for observed liver tumors in
animals. Rosen et al. (2008a, 2008b) examined transcript profiles in the livers of wild-type and
PPARa-null mice dosed with PFOA for up to 7 days. This study showed that animal responses
were consistent with PPARa agonism, but evidence also shows PPARy agonism (down-
regulation of cholesterol synthesis) and activation of CAR and PXR-related genes (Martin et al.
2007). There is evidence that PFOA is a potent peroxisome proliferator inducing peroxisome
formation in the livers of rats and mice (Elcombe et al. 2010; Minata et al. 2010; Pastoor et al.
1987; Wolf et al. 2008; Yang et al. 2001). Beyond activation of PPARa, few studies have
evaluated whether additional steps (i.e., cell proliferation and apoptosis) are in the hypothesized
PPARa agonism MOA (Elcombe et al. 2010; Minata et al. 2010; Wolf et al. 2008). For example,
no studies were identified that focused specifically on preneoplastic foci and clonal expansion of
altered cells after PPAR activation.
The proposed MOA for testicular LCTs is linked to decreased serum testosterone levels and
signaling of the hypothalamus to produce gonadotropin releasing hormone (GnRH), a signaling
agent for the pituitary to release luteinizing hormone which upregulates testosterone production
in Leydig cells. Administering PFOA to adult male rats by gavage for 14 days was shown to
decrease testosterone levels and increase serum estradiol levels (Cook et al. 1992). These
endocrine changes correlated with its potency to induce LCTs in rats and were hypothesized to
play a role in the PFOA-induction of LCTs (Biegel et al. 2001). Support for PPARa-mediated
inhibition of testosterone production is found in Li et al. (2011). However, some researchers
have proposed that data are not currently sufficient to demonstrate that the other key steps in the
postulated MOA are present in PFOA-treated animals following exposures that lead to tumor
formation (Klaunig et al. 2012).
Two hypothetical MO As have been proposed for PACTs (Klaunig et al. 2003, 2012; Obourn
et al. 1997). In one case, growth factors such as cholecystokinin (CCK) and/or gastrin activate a
feedback loop resulting in proliferation of the secretory pancreatic acinar cells leading to tumors.
The other proposed MOA suggests that increased serum testosterone supports the growth of
acinar cell preneoplastic foci.
Li et al. (2011) found that serum testosterone levels were decreased, not increased, in wild-
type, PPARa- null and mice with humanized PPARa. Biegel et al. (2001) found no change in
serum testosterone in their bioassay. Obourn et al. (1997) studied the impact of PFOA on CCK
and trypsin using in vitro assays and found that PFOA was not an agonist for the cholecystokinin
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) - May 2016 46
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agonism receptor receptor that activated CCK release. Plummer et al. (2007) reported on gene
expression changes induced in pancreatic acinar cells isolated from Sprague-Dawley rats fed
diets containing 300 parts per million (-20 mg/kg/day) PFOA for 28 days. Expression of genes
regulated by PPARa, y, 5 in pancreatic acinar cells was directly opposite of the expression of
those same genes in liver tissue. At the present time, data are insufficient to demonstrate a MOA
that can account for the PACTs identified in the chronic study by Biegel et al. (2001).
The mutagenicity data on PFOA are largely negative, although some evidence shows
clastogenicity in the presence of microsomal activation and at cytotoxic concentrations (Murli
1996a, 1996b). PFOA's clastogenic effects are likely the result of an indirect mechanism, given
the chemical and physical properties of PFOA (i.e., it is not metabolized, it binds to cellular
proteins, and it carries a net negative electrostatic surface charge). PFOA has the potential to
interfere with the process of DNA replication because of its protein-binding properties and the
fact that histone proteins, spermine, and spermidine carry a net positive surface charge.
Involvement of reactive oxygen species (ROS) in the MOA as a result of PFOA alone is unlikely
because of its metabolic stability. Conditions leading to ROS would be a function of metabolic
responses perturbed by PFOA, rather than PFOA alone.
A compound that is not metabolized will not be able to covalently alter the structure of DNA
or intercalate because of electrostatic repulsion between the aromatic base pi bond electrons and
the partial negative charges on the PFOA fluoride atoms. Because of its protein-binding
properties, PFOA could affect one or more of the proteins involved in the process of DNA
replication or cell division (cytoskeletal proteins), however, no mechanistic studies were
identified that examined the biochemical effects of PFOA on DNA replication or cell division.
No data support this as a MOA for clastogenic effects.
4.2.4 Weight of Evidence Classification
Under EPA's Guidelines for Carcinogen Risk Assessment (USEPA 2005) there is Suggestive
Evidence of Carcinogenic Potential of PFOA in humans. The bioassay findings for Ley dig cell
testicular tumors in rats, combined with the C-8 Panel finding of a probable link to testicular and
renal tumors among the members of the C8 Health Project, support this conclusion.
In June 2014, 20 experts met at the International Agency for Research on Cancer (IARC) in
Lyon, France, to assess the carcinogenicity of perfluorooctanoic acid (PFOA), among other
chemicals. Although the assessments have not yet been published (they are expected to be
published in volume 110 of the IARC monographs), the expert findings from this meeting are
available in a peer-reviewed publication (Benbrahim-Tallaa et al. 2014), and their determination
is on the IARC website. The working group classified PFOA as possibly carcinogenic to humans
(Group 2E) and considered the evidence regarding mechanisms of PFOA-associated
carcinogenesis to be moderate. This assessment did not lead to a change in the overall
classification of PFOA by IARC.
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5.0 DOSE-RESPONSE ASSESSMENT
As an initial step in the dose-response assessment, EPA identified a suite of animal studies
with NOAELs and/or LOAELs that identified them as potential candidates for development of
the RfD for PFOA. These studies included short-term, subchronic, and chronic exposures,
including developmental and reproductive toxicity studies. The available studies evaluated
endpoints including liver effects (weight changes with histopathology), body weight changes in
adults and offspring, reproductive outcomes such as fertility, developmental effects (altered
puberty, survival, and developmental delays such as eye opening), and immune effects. The
candidate studies were selected based on their NOAEL and/or LOAEL values, a duration of
11 to 91 days, use of a control, and two or more doses. From these studies, those that presented
serum data amenable for modeling (i.e., determination of HEDs) were selected for dose-response
analysis. The subset of studies amenable for use in deriving FED based on average serum
measurements from the pharmacokinetic model is limited because of the need to have dose and
species-specific serum values for model input as well as exposure durations of sufficient length
to achieve values near to steady-state projections or applicable to developmental endpoints with
lifetime consequences following short-term exposures. The pharmacokinetically modeled
average serum values from the animal studies are restricted to the animal species selected for
their low dose response to oral PFOA intakes.
As described in section 3.2.4, EPA used the Wambaugh et al. (2013) pharmacokinetic model
to derive the average serum concentrations associated with the candidate NOAELs and LOAELs
from the toxicological database. Studies with serum information for each of the doses that
demonstrated dose response and were amendable for modeling of the area under the curve
(AUC) at the time of sacrifice were used. The AUC results were converted to average serum
values at the time of sacrifice with consideration of the duration of exposure. The average serum
values were converted to the FLED, as described further below.
The data were analyzed within a Bayesian framework using a Markov Chain Monte Carlo
sampler implemented as an R statistical analysis package developed by EPA to allow predictions
across species, strains, and genders, and to identify serum levels associated with the external
doses at the NOAEL and LOAEL. The model predictions were evaluated by comparing each
predicted final serum concentration to the serum value measured in the supporting animal
studies.
The average serum concentrations were converted into an oral equivalent dose by
recognizing that clearance from the body equals dose to the body. Clearance can be calculated if
the rate of elimination (derived from half-life) and the volume of distribution are both known.
EPA used the Bartell et al. (2010) calculated human half-life of 2.3 years (general population)
with the Thompson et al. (2010) volume of distribution (Vd) of 0.17 L/kg body weight (bw) to
determine a clearance of 1.4 x 10~4L/kg bw/day by the following equation:
CL = Vd x (In 2 H- t%) = 0.17 L/kg bw x (0.693 H- 839.5 days) = 0.00014 L/kg bw/day
Where:
Vd = 0.17L/kg
In 2 = 0.693
tvi = 839.5 days (2.3 years x 365 days/year = 839.5 days)
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Multiplying the derived average serum concentrations (in |ig/mL) for the NOAELs and
LOAELs identified in the key animal studies by the clearance value predicts oral HEDs in mg/kg
bw/day for each corresponding serum measurement. The HED values are the predicted human
oral exposures necessary to achieve serum concentrations equivalent to the NOAEL or LOAEL
in the animal toxicity studies using linear human kinetic information.
Table 5-1 provides the NOAEL, LOAEL, and effect information from those studies, along
with the associated average serum values and the percent of steady state represented by the
LOAEL.
Table 5-1. Human Equivalent Doses Derived from the Modeled Animal
Average Serum Values
Study
De Witt etal. (2008):
mice; J, IgM response
toSRBC
Lauetal. (2006): mice
decreased J, pup
ossification (m, f),
accelerated male
puberty
Palazzolo etal. (1993);
Perkins etal. (2004):
rats; fliver
weight/necrosis
Wolf etal. (2007):
mice; GD 1-17
|Pup body weight
Wolf etal. (2007):
mice; GD 7-17
|Pup body weight1
Butenhoff et al.
(2004a): | relative
body weight/t relative
kidney weight and
fkidney: brain weight
ratio inFO andFl at
sacrifice
Dosing
duration
days
15
17
91
17
11
84
NOAEL
mg/kg/d
1.88
None
0.64
None
None
None
NOAEL
Av serum
mg/L
38.2
-
31.6
-
-
-
HED
mg/kg/d
0.0053
-
0.0044
-
-
-
LOAEL
mg/kg/d
3.75
1
1.94
o
J
5
1
LOAEL
(Av serum)
mg/L
61.9
38.0
77.4
77.9
87.9
45.9
HED
mg/kg/d
0.0087
0.0053
0.0108
0.0109
0.0123
0.0064
Notes'.
Significance p < 0.05 or p < 0.01
m = male; f = female; SRBC = sheep red blood cell; IgM = immunoglobulin M; GD = gestation day
1 serum from pups on PND 22
The external doses in each of the studies varied. The NOAELs ranged from 0.64 to
1.88 mg/kg/day. The corresponding average serum values ranged from 1.6 mg/L (rat) to
38.2 mg/L (mouse). At the LOAEL, the average serum values range from 38 |ig/mL (mouse) to
87.6 |ig/mL (monkey) at doses estimated to represent about 56% to 96 % of steady state. At the
low end of the range the effects of concern are observed in neonates (low birth weight, delays in
developmental endpoints, with increased kidney weight at sacrifice later in life).
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Much of the variability in the average serum levels for the LOAELs was due to differences in
the doses used in the individual studies. For example, two of the modeled endpoints (Wolf et al.
2007) identified low birth weights in mouse pups as the critical effect, but had a single external
dose that was 3 to 5 times higher than the low dose from the Lau et al. (2006) mouse study
(1 mg/kg/day).
Among the studies conducted in mice, dose was a more important variable in determining
serum level and percent of steady state than duration of exposure. This is a characteristic of the
nonlinear toxicokinetics exhibited by PFOA. The half-life for doses that exceed the resorption
capacity of the kidney are shorter than lower doses that can be resorbed and thereby persist in
serum over a longer exposure duration. For example, in Wolf et al. (2007), an 11-day dose of
5 mg/kg/day resulted in an average serum of 88 mg/L (82% of concentration at steady state or
Css) whereas a 1 mg/kg/day dose for 17 days resulted in an average serum of 38 mg/L (56% of
Css). In rats, dosed at 1 mg/kg/day, over two generations (84 days), an average serum of
45.9 mg/L at 87% of steady state was determined (Butenhoff et al. 2004a). A 91-day exposure
(Palazzolo et al. 1993/Perkins et al. 2004) to 1.94 mg/kg/day resulted in a serum value of
77 mg/kg/day and was 91% of steady state. The endpoints in Butenhoff et al. (2004a) are effects
on body weight and relative kidney weight in the adult FO and Fl rats, while the endpoint for
Palazzolo et al. (1993)/Perkins et al. (2004) was systemic increased liver weight with lower-level
necrosis.
Assuming that MOA and susceptibility to toxicity do not vary and that pharmacokinetics
alone explains variation, it is reasonable to expect similar concentrations to cause similar effects
in humans and are more important than both dose and duration once steady state is attained.
5.1 Uncertainty Factors
An uncertainty factor for intraspecies variability (UFn) of 10 is assigned to account for
variability in the responses within the human populations because of both intrinsic (toxicokinetic
genetic, life stage, health status) and extrinsic (life style) factors that can influence the response
to dose. No information was available relative to variability in the human population that
supports a factor other than 10.
An uncertainty factor for interspecies variability (UFA) of 3 is applied to account for
uncertainty in extrapolating from laboratory animals to humans (i.e., interspecies variability).
The 3-fold factor is applied to account for toxicodynamic differences between the animals and
humans. The HEDs were derived using average serum values from a model to account for
toxicokinetic differences between animals and humans.
An uncertainty factor for LOAEL to NOAEL extrapolation (UFL) of 10 is applied to all
PODs other than the Palazzolo et al. (1993)/Perkins et al. (2004) and DeWitt et al. (2008) studies
to account for use of a LOAEL for the POD. The POD for the Palazzolo et al. (1993)/Perkins et
al. (2004) and DeWitt et al. (2008) studies are NOAELs for the effect identified as critical.
An uncertainty factor for extrapolation from a subchronic to a chronic exposure duration
(UFs) of 1 is applied because the PODs are based on average serum concentrations and
determined to represent >80% of steady state for each study (81-91%), except for the Lau et al.
(2006) developmental study (56%). The Lau et al. (2006) developmental FLED was not adjusted
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) - May 2016 50
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for lifetime exposures because the average serum values associated with the developmental
studies are more protective than those for the longer-term studies of systemic toxicity. A UFs of
10 was applied to the DeWitt et al. (2008) study serum derived HED reflecting (74%) of steady
state because the data suggest that longer term exposures to the same dose have the potential to
increase serum values beyond the levels indicated by the 15-day study. In addition, the NOAEL
for immunological effects (0.94 mg/kg/day) was a LOAEL for effects on liver weight in the
absence of histological evaluation on both days 16 and 31 following a 15-day exposure (DeWitt
et al. 2008). Thus, there is a potential that lifetime exposures at steady state can affect the liver
and increase the risk for tissue damage.
A database uncertainty factor (UFo) of 1 was applied to account for deficiencies in the
database for PFOA. There are extensive human data from epidemiological data from the general
population as well as worker cohorts. The epidemiology data provide strong support for the
identification of hazards observed following exposure to PFOA in the laboratory animal studies
and human relevance. However, uncertainties in the use of the available epidemiology data
precluded their use at this time in the quantification of the effect level for derivation of the
drinking water HA. In animals, acute, short term, subchronic and chronic studies, including a
long term cancer study, are available. In addition, several developmental studies and a two-
generation reproductive toxicity study evaluating exposure of pregnant dams and offspring to
PFOA are available.
5.2 RfD Determination
Table 5-2 provides the calculations for candidate RfDs using the HEDs derived from the
NOAEL or LOAEL average serum concentrations using pharmacokinetic modeling based on the
serum values measures collected at animal sacrifice. Uncertainty factors (see section 5.1) were
applied to each POD, and Table 5-2 illustrates the array of candidate RfD outcomes. Each POD
is affected by the doses used in the subject study, the endpoints monitored, and the animal
species/gender studied. Thus, the array of outcomes, combined with knowledge of the individual
study characteristics helps to inform selection of an RfD that will be protective for humans.
Other than DeWitt et al. (2008) and Lau et al. (2006), all of the selected studies had serum levels
that had reached > 80% of Css. It is important to note the relatively narrow range of RfDs across
the multiple endpoints and study durations evaluated.
Using the pharmacokinetic model of Wambaugh et al. (2013), average serum PFOA
concentrations were derived from AUC considering the number of days of exposure before
sacrifice. The predicted serum concentrations were converted as described above to oral HEDs
mg/kg/day for each corresponding serum measurement. The candidate RfDs in Table 5-2 range
from 0.00002 to 0.00015 mg/kg/day across multiple endpoints. The RfD of 0.00002 mg/kg/day
calculated from HED average serum values from Lau et al. (2006) was selected. This RfD is
derived from reduced ossification of the proximal phalanges (forelimb and hindlimb) and
accelerated puberty in male pups (4 days earlier than controls) as the critical effects. The POD
for the derivation of the RfD for PFOA is the HED of 0.0053 mg/kg/day that corresponds to a
LOAEL that represents approximately 60% of steady-state concentration. An UF of 300
(10 UFn, 3 UFA, and 10 UFL) was applied to the HED LOAEL to derive an RfD of 0.00002
mg/kg/day.
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) - May 2016 51
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Table 5-2. Candidate RfDs Derived from the HEDs from the Pharmacokinetic Model
Average Serum Values
POD
PK-HEDNoAEL Palazzolo
etal. (1993)/Perkinsetal.
(2004)
rats; fliver
weight/necrosis
PK-HEDLOAEL Wolf etal.
(2007) GD1-17 mice;
|Pup body weight
PK-HEDLOAEL Wolf etal.
(2007) GD 7-17 mice;
|Pup body weight (serum
from pups on PND 22)
PK-HEDNOAEL DeWitt et
al. (2008)
mice; J, IgM response to
SRBC
PK-HEDLOAEL Lau etal.
(2006)
mice decreased J, pup
ossification (m, f),
accelerated male puberty
PK-HEDLOAEL Butenhoff
et al. (2004a)
I relative body weight/t
relative kidney weight
and fkidney: brain weight
ratio inFO andFl at
sacrifice
RED POD
mg/kg/day
0.0044
0.0109
0.0123
0.0053
0.0053
0.0064
UFH
10
10
10
10
10
10
UFA
o
J
o
J
3
3
o
5
o
6
UFL
10
10
10
10
UFs
10
UFD
UFtotal
30
300
300
300
300
300
Candidate RfD
mg/kg/day
0.00015
0.00004
0.00004
0.00002
0.00002
0.00002
Notes'.
PK-HED = pharmacokinetic human equivalent dose; NOAEL = no observed adverse effect level; LOAEL = lowest observed
adverse effect level; GD = gestation day; IgM = immunoglobulin M; m = male; f = female; SRBC = sheep red blood cell; UFn =
intraindividual uncertainty factor; UFA = interspecies uncertainty factor; UFs = subchronic to chronic uncertainty factor; UFL =
LOAEL to NOAEL uncertainty factor; UFo = incomplete database uncertainty factor; UFtotal = total (multiplied) uncertainty
factor
Decreased pup body weights also were observed in studies conducted by Wolf et al. (2007),
White et al. (2009), and Lu et al. (2015) using mice receiving external doses within the same
order of magnitude (1,3, and 5 mg/kg/day respectively) as those chosen for the RfD. The
selected RfD from the reproductive and developmental studies is supported by the longer term
RfD for effects on the response of the immune system to external challenges as observed
following the short-term exposures to mature mice and the effects on kidney weight observed at
the time of sacrifice in the FO and F1 adult males that provided the serum in the Butenhoff et al.
(2004a) study (DeWitt et al. 2008).
Support for the selected RfD also is provided by other key studies with NOAELs and
LOAELs similar to those used for quantification, but lacking serum data that could be used for
modeling. There were effects on liver weight and hepatic hypertrophy in the Perkins et al. (2004)
and DeWitt et al. (2008) studies that were modeled but not considered in the derivation of the
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RfD because of a lack of data to demonstrate adversity as determined by the Hall et al. (2012)
criteria at the dose causing the liver effects but not the effects identified as critical. The LOAEL
for evidence of hepatic necrosis and other signs of tissue damage in the Fl male rat pups from
the Butenhoff et al. (2004a) study was 3 mg/kg/day; the NOAEL was 1 mg/kg/day. In the
Loveless et al. (2008) study, the LOAEL for increased relative liver weight accompanied by
focal liver necrosis in male rats was 10 mg/kg/day and the NOAEL was 1 mg/kg/day, while in
male mice, the LOAEL for the same effect was 1 mg/kg/day and the NOAEL was 0.3 mg/kg/day
following a 29-day exposure. In the study by Tan et al. (2013), the degree of damage to the liver
at 5 mg/kg/day became more severe with increased necrosis, inflammation, and steatosis when
animals were given a high-fat diet. The HED modeled from the average serum value in mice for
the LOAEL (3 mg/L) from Wolf et al. (2007) and White et al. (2009) was 0.0110 mg/kg/day,
about twice that for the rats in the Lau et al. (2006) study (0.0053 mg/kg/day). Both studies
lacked a NOAEL. Each of these data sets support LOAELs for the critical study by Lau et al.
(2006) selected for RfD derivation and, as a consequence, the HED derived from modeled
average serum values.
6.0 HEALTH ADVISORY VALUES
6.1 Relative Source Contribution
As described in section 2.2 and below, humans can be exposed to PFOA and precursor
chemicals via multiple sources, including air, food, and consumer and industrial products
(including textiles and rugs). The most common route of exposure to PFOA is via the diet,
followed by indoor dust, especially for children.
Food is a significant source of exposure to PFOA: It has been detected in a variety of foods
including snack foods, vegetables, meat, dairy products, human breast milk, and fish. Occurrence
in food products can result from the use of contaminated water in processing and preparation;
growth of food in contaminated soils; direct and indirect exposures of domestic animals to PFOA
from drinking water, consumption of plants grown in contaminated soil, and through particulate
matter in air; fish from contaminated water ways; and packaging materials.
PFOA has been detected in finished drinking water samples collected by EPA and others.
PFOA is not regulated under the SDWA and was included in EPA's UCMR 3. PFOA was
detected at a small number of PWSs (0.9%) through this monitoring program. Therefore, there is
potential exposure to PFOA from drinking water ingestion.
The vapor pressure of PFOA indicates that volatilization is low; however, PFOA can be
released into the atmosphere from industrial and municipal waste incinerators and adsorb to
airborne particulates. It can be transported long distances through the atmosphere and has been
detected globally at low concentrations. Inhalation of PFOA is possible, and it has been
measured in indoor air in residential, commercial, and office settings because of its use in
carpets, textiles, paint, furniture, and other consumer products. Both air and dust can be a vehicle
for volatile telomer alcohols that metabolically degrade to PFOA. Given the widespread
commercial and industrial use of PFOA and its physical properties, air is a potential source of
exposure to it and the C8:2 telomer alcohol precursors.
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PFOA also has been detected in soils and dust from carpets and upholstered furniture in
homes, offices, and vehicles. Incidental exposure from soils and dust is an important exposure
route, particularly for small children because of their hand-to-mouth behaviors. Also, the levels
in soils and surface waters can affect the concentrations in local produce, meat/poultry, dairy
products, fish, and particulates in the air.
In summary, based on the physical properties and available exposure information for PFOA,
there are many are potential sources. Following EPA's Exposure Decision Tree in its 2000
methodology (USEPA 2000), significant potential sources other than drinking water ingestion
exist; however, information is not available to quantitatively characterize exposure from all of
these different sources (Box 8B in the Decision Tree). Therefore, EPA recommends an RSC of
20% (0.20) for PFOA.
6.2 Lifetime Health Advisory
Based on the consistency of the responses across the chronic studies and those for
reproductive and developmental endpoints, and with recognition of the use of developmental
toxicity as the most sensitive endpoint, 0.00002 mg/kg/day was selected as the RfD for PFOA.
This value is based on the HED for developmental effects (reduced ossification in male and
female pups and accelerated puberty in male pups) from the Lau et al. (2006) study. The RfD
that serves as the POD for the lifetime HA is applicable for effects other than those occurring
during development. The candidate RfD values derived from the two-generation study by
Butenhoff et al. (2004a) for effects on adult body weight plus relative liver and kidney weights in
FO and Fl male rats is the same as the value based on the developmental effects observed by Lau
et al (2006). The candidate RfD from the DeWitt et al. (2008) study for suppression of the
immunological response to a challenge is the same as that from Lau et al. (2006).
Due to the potential increased susceptibility during the time period of pregnancy and
lactation, EPA used drinking water intake and body weight parameters for lactating women in
the calculation of a lifetime HA for this target population during this potential critical time
period. EPA used the rate of 54 mL/kg-day representing the consumers only estimate of
combined direct and indirect community water ingestion at the 90th percentile for lactating
women (see Table 3-81 in USEPA 201 Ib). Comparing the pregnant woman and the lactating
woman, the lactating woman is the more protective scenario given her increased water intake
rate for her body weight needed to support milk production. Additionally, human studies
demonstrate that PFOA is transferred from mother to infant via cord blood and breast milk. A
recent study showed that breast milk contributed > 83% of the PFOA exposure in 6-month-old
infants (Haug et al. 2011).
The exposure factors applied to the RfD to derive the lifetime HA are specific to the most
sensitive population and will be protective of pregnant women as well as of the general
population. Thus, the protection conferred by the lifetime HA is broadly protective of public
health.
The lifetime HA for PFOA is calculated as follows:
A DWEL is derived from the RfD and assumes that 100% of the exposure comes from
drinking water.
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RfD x bw
DWEL = 0.00002 mg/kg/day = 0.00037 mg/L
0.054 L/kg-day
Where:
RfD = 0.00002 mg/kg/day; based on the LOAEL for reduced ossification of the proximal
phalanges (forelimb and hindlimb) in male and female pups and accelerated (4 days
earlier than controls) puberty in male pups of dams exposed to PFOA by gavage on
gestation days 1 to 17 and sacrificed at weaning (Lau et al. 2006).
DWI/bw = 0.054 L/kg-day; 90th percentile consumers only estimate of combined direct
and indirect community water ingestion for lactating women (see Table 3-81 in
USEPA2011b).
The lifetime HA is calculated after application of a 20% RSC (see section 6.1) as follows:
Lifetime HA = DWEL x RSC
= 0.00037 mg/L x 0.2
= 0.000074 mg/L (rounded to 0.00007 mg/L)
= 0.07 |ig/L
The lifetime HA for PFOA is based on effects (reduced ossification in male and female pups
and accelerated puberty in male pups) on the developing fetus resulting from exposures that
occur during gestation and lactation. These developmental endpoints are the most protective for
the population at large and are effects that can carry lifetime consequences for a less than
lifetime exposure. Developmental toxicity endpoints (following less than chronic exposures
during a defined period of gestation or lactation) can be analyzed in both acute and chronic
exposure scenarios. Because the developing organism is changing rapidly and is vulnerable at
various stages in development, a single exposure at a critical time in development can produce an
adverse effect (USEPA 1991). Additionally, PFOA is extremely persistent in both the human
body and the environment; thus, even a short-term exposure results in a body burden that persists
for years and can increase with additional exposures.
Because the critical effect identified for PFOA is a developmental endpoint and can
potentially result from a short-term exposure during a critical period of development, EPA
concludes that the lifetime HA for PFOA is applicable to both short-term and chronic risk
assessment scenarios. Thus, the lifetime HA of 0.07 ug/L also applies to short-term exposure
scenarios (weeks to months) to PFOA in drinking water, including during pregnancy and
lactation.
Adverse effects observed following exposures to PFOA and PFOS are the same or similar
and include effects on serum lipids, birth weight, and serum antibodies in humans. Among the
animal studies, there are common effects on the liver, neonate development, and responses to
immunological challenges. Both compounds also were associated with tumors in long-term
animal studies. The effects that serve as the basis for the RfDs for both PFOA and PFOS are
developmental endpoints (reduced ossification and accelerated puberty in males for PFOA and
decreased pup birth weight for PFOS (USEPA 2016a, 2016b). Because the RfDs for both PFOA
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and PFOS are based on similar developmental effects and are numerically identical, where these
two chemicals co-occur at the same time and location in a drinking water source, a conservative
and health protective approach that EPA recommends would be to compare the sum of the
concentrations ([PFOA] + [PFOS]) to the HA (0.07 ug/L).
7.0 QUANTIFICATION OF CANCER RISK
The evidence for the carcinogenicity of PFOA is considered suggestive because only one
species has been evaluated, and the tumor responses (liver, testicular Leydig cell, and pancreatic
acinar cell tumors) occurred primarily in males. Dose-response data are only available for the
LCTs in one study. The dose-response data on LCTs from the (Butenhoff et al. (2012) studies
were modeled to provide a perspective on the magnitude of the potential cancer risk as it
compares with the level of protection provided by the RfD.
Under EPA's Guidelines for Carcinogen Risk Assessment (USEPA 2005), when there is
Suggestive Evidence for Carcinogenic Potential for a chemical, a dose-response assessment
would generally not be attempted. The guidelines state that, when the evidence includes a well-
conducted study, quantitative analyses could be useful for some purposes—for example, by
providing a sense of the magnitude and uncertainty of potential risks, ranking potential hazards,
or setting research priorities. The data from the Butenhoff et al. (2012) study are adequate to
support a quantitative cancer dose-response assessment for PFOA's testicular tumors. The
epidemiology studies demonstrate an association of serum PFOA with kidney and testicular
tumors among highly exposed members of the general population. Thus, EPA concluded that a
quantitative analysis could be useful by providing a sense of the magnitude of potential
carcinogenic risk.
The dose-response data for LCTs in rats was analyzed using the multistage cancer model for
a dichotomous data set to predict the dose at which a 4% increase in tumor incidence would
occur (see appendix A). A benchmark response of 4% was chosen as the low end of the observed
response range within the study results. The resulting benchmark dose level (BMDLo4) was
1.99 mg/kg/day, which yields a FLED of 0.58 mg/kg/day and a slope factor of 0.07 (mg/kg/day)"1.
The cancer slope factor was calculated to determine if a lifetime HA derived from the RfD would
be protective for the cancer endpoint. As a comparative analysis, the concentration of PFOA in
drinking water that would have a one-in-1-million chance of an increased cancer risk was
calculated using the oral slope factor for testicular tumors, assuming a default adult body weight
of 80 kg (mean weight for adults ages 21 and older) (Table 8.1 in USEPA 201 Ib) and a default
drinking water intake rate of 2.5 L/day (consumers only estimate of combined direct and indirect
community water ingestion at the 90th percentile for adults ages 21 and older) (Table 3-33 in
USEPA 201 Ib). The resultant 0.5 ug/L value is greater than the lifetime HA (0.07 ug/L) based
on noncancer effects (see section 6.22.2.), indicating that the HA derived based on the
developmental endpoint is protective for the cancer endpoint.
10-6 cancer risk = 0.000001 x 80 kg = 0.00046 mg/L rounded to 0.5 ug/L
(0.07 mg/kg/day x 2.5L/d)
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8.0 EFFECTS CHARACTERIZATION
8.1 Uncertainty and Variability
The variability and uncertainty in the lifetime HA is a function of both intrinsic and extrinsic
factors. EPA's HESD for PFOA identified 20 short- or long-term studies that provided dose-
response information and were considered during the quantitative assessment of risk (USEPA
2016a). The range of external dose NOAELs among the 20 studies is 0-30 mg/kg/day for
females and 7.5 mg/kg/day for males. The LOAELs range from zero to 30 mg/kg/day for
females and zero to 14.2 mg/kg/day for males (USEPA 2016a). Six dose-response data sets
included the serum data necessary for modeling to derive HEDs for use as the POD for the RfD.
Average serum values from those studies were chosen for use in the derivation of the RfD. The
external dose range for the NOAELs in the modeled studies is 0-1.88 mg/kg/day and the
LOAEL range is 1-5 mg/kg/day (USEPA 2016a). EPA believes the uncertainty in the chosen
POD and the reliance on studies with serum data is minimized because of the large and extensive
database examining the PFOA hazard and the selection of reduced ossification, and accelerated
male puberty as the critical effects with lifetime implications at a LOAEL dose (1 mg/kg/day)
from the low end of the narrow range of values evaluated.
The intrinsic uncertainties in the risk assessment reflect the fact that the NOAELs and
LOAELs are derived using central tendency estimates for variables such as body weight, food
and drinking water intakes, and dose. The central tendency estimates are derived from small
numbers of genetically, relatively similar animals representing one or more strains of rats or
mice living in controlled environments. The animals lack the heterogeneous genetic complexity,
behavioral diversity, and complex habitats experienced by humans. These differences, to some
extent, are minimized through the use of the modeled central tendency outcomes and their
standard deviations to help inform the application of the uncertainty factors.
Variability in the study outcomes is extrinsically a function of study design and the endpoints
monitored. Studies of systemic toxicity monitor an array of endpoints that are not evaluated in
studies of reproductive, developmental, neurological, and immunological toxicity. The reverse is
true for the other types of toxicity studies compared to standard short-term to long-term systemic
studies. Studies of systemic toxicity do not often examine neurological or immunological
endpoints. Increases in liver weight were seen in many of the studies with dose-response
information but only a few of the studies carried out a histological evaluation of the liver to
support a determination of whether the increase in liver weight could be classified as adverse
according to the Hall et al. (2012) criteria.
The RfD is based on the HED derived from serum levels at the LOAEL from developmental
study in mice with application of an uncertainty factor of 300 to cover extrapolation from a
LOAEL to a NOAEL, variability in the human population, and differences in the ways humans
and rodents respond to the PFOA that reaches their tissues (Lau et al. 2006). The selected RfD is
based on the developmental effects in neonates to provide protection to both the sensitive life
stages and the general population. The RfD is supported by the outcomes from two other studies
based on different endpoints, Butenhoff et al. (2004a) and DeWitt et al. (2008), with RfDs for
systemic effects on liver, kidney and the immune system. These data increase the confidence in
the RfD.
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8.2 Use of Human Epidemiology Data
The human epidemiology studies provide evidence of an association between PFOA
exposure and health effects in humans, and is another line of evidence supporting this
assessment. The human data demonstrate an association between PFOA exposure and endpoints,
including effects on serum lipids, antibody responses, fetal growth and development, and the
liver. They provide support for identification of hazards of PFOA exposure. The associations
observed for serum lipids, and reproductive parameters and immunotoxicity are the strongest.
For many endpoints, however, the results are inconsistent. Although the human studies
collectively support the conclusion that PFOA exposure is a hazard, EPA concluded that, based
on several uncertainties associated with the database, the human studies are adequate for use
qualitatively in the identification hazard at this time. These considerations are discussed below.
Although mean serum values are presented in the human studies, actual estimates of
exposure (i.e., doses/duration) are not available. Thus, the serum level at which the effects were
first manifest and whether the serum had achieved steady state or was in decline at the point the
effect was evaluated cannot be determined. The NHANES and C8 study data indicate that serum
levels in the general population are declining. Because epidemiology data are a reflection of the
serum concentration at the time the sample was collected, there is no way to determine if levels
were previously higher and had decreased. All of the C8 study serum samples were collected
after the PFOA peak exposures had presumably passed. The half-life measurement for the
general population of the Little Hocking area was derived from declining serum concentrations
over time, demonstrating that serum levels among that population were not constant (Bartell et
al. 2010).
Some of the human exposure that results in serum PFOA can come from telomer alcohol
PFOA derivatives that break down metabolically to PFOA (Gebbink et al. 2015; Jogsten et al.
2012). The derivatives do not originate from PFOA in drinking water; they usually originate
from diet and materials used in the home. Thus, there is added uncertainty in the observed
epidemiological associations between serum PFOA and health effects.
Although the epidemiology studies provide valuable associations between exposure to PFOA
and the effects seen in animal studies, most of the subjects of the epidemiology studies had other
perfluorinated carboxylates and sulfonates and/or other biopersistent contaminants in their blood.
Although the study designs adjust for other potential toxicants as confounding factors, their
presence constitutes a level of uncertainty that is usually absent in the animal studies.
The database for PFOA includes extensive human data from epidemiology studies of the
general population as well as worker cohorts. Data from oral short-term, subchronic, chronic
(including evaluation of cancer), reproductive, and developmental studies in laboratory animals
also are available. Many of the effects observed in the human epidemiology studies are similar to
those seen in the animal studies.
8.3 Consideration of Immunotoxicity
Both human and animal studies have demonstrated the potential effect of PFOA on the
immune system. However, there are uncertainties related to MOA and the level, duration, and/or
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timing of exposure that are not yet clearly delineated. As a result, EPA used the animal data
rather than the human data to quantify the dose response for immunotoxicity for PFOA.
Taken together, available human studies do not provide consistent evidence of a significant
association between PFOA exposure and serological vaccine responses in general (Grandjean et
al. 2012; Granum et al. 2013; Looker et al. 2014). Within each study, most estimated
associations were statistically nonsignificant, and results were inconsistent by vaccine type and
by outcome classification. Authors provided no a priori biological hypothesis to explain why
PFOA exposure would impair the antibody response to one vaccine type but not another. Some
authors suggested that their results could be explained by different immunostimulatory effects of
different vaccines, but they did not elaborate on this hypothesis or provide supporting
mechanistic evidence.
One issue related to use of immune biomarkers and antibody levels in human studies is
whether small but statistically significant changes in these endpoints, when analyzed on a
continuous scale, are clinically meaningful, particularly when most or all subjects are within the
normal range. For PFOA, some epidemiology studies attempted to address this issue by
analyzing outcomes dichotomized relative to standard reference values, with the implication that
values outside the reference range indicate immune abnormalities (Emmett et al. 2006;
Grandjean et al. 2012; Looker et al. 2014). A limitation of this approach is that a reference range
is typically determined based on the mean plus or minus two standard deviations calculated from
a group of healthy adults or children. By definition, 5% of the normal population falls outside of
such a reference range (AACC 2015, cited in Chang et al. 2016). The only way to determine
whether a given value outside a reference range is truly abnormal is to associate it with a clinical
abnormality; this has not been done in most epidemiology studies of immune biomarkers.
Although Grandjean et al. (2012) found fairly consistent, albeit mostly statistically
nonsignificant, intrastudy associations between childhood serum PFOA levels and poorer
antibody responses against tetanus and diphtheria toxoids, associations with maternal prenatal
serum PFOA and PFOS levels were inconsistent between vaccine types. Two studies were
strengthened by their measurement of PFOA levels prior to ascertaining vaccine response
(Grandjean et al. 2012; Granum et al. 2013), and one had the additional advantage of collecting
exposure and outcome information at two time points each (Grandjean et al. 2012). However, the
variability in findings by timing of exposure and outcome measurement in the latter study
(e.g., mostly nonsignificant associations with prenatal PFOA concentrations, but several
significant associations between higher PFOA concentrations at age 5 years and poorer vaccine
response at age 7 years) makes the results difficult to interpret. This pattern of results could
reflect a window of susceptibility in early childhood, but such an explanation remains
conjectural.
None of the studies demonstrated a clinically recognizable increased risk of infectious
diseases as a consequence of a diminished vaccine response. Overall, although these results are
not sufficient to establish a causal effect of PFOA exposure on an impaired serological vaccine
response, some of the positive associations are striking in magnitude and require replication in
independent studies.
Chang et al. (2016) recently completed and published a systematic review of 24
epidemiology studies that reviewed a variety of endpoints among the general population,
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occupationally exposed workers, children, and adults and concluded that the available
epidemiologic evidence is insufficient to reach a conclusion about a causal relationship between
exposure to PFOA and PFOS and any immune-related, health condition in humans. The majority
of the studies reviewed by the authors are included in EPA's FIESDs for PFOA and PFOS
(USEPA 2016a, 2016b). The authors identified numerous weaknesses in study designs, including
lack of validation of self-reported medical conditions, basing conclusions on significant
associations without considering statistical significance, inadequate consideration of
confounding factors, bias, and the role of chance being responsible for outcomes. After
application of the Hill et al. (1965) criteria, they faulted the studies for "generally weak
associations, no specific endpoints with consistent findings across all relevant studies,
uncertainty about any critical duration of exposure and window(s) of susceptibility, mixed
exposure-response trends, and a dearth of supportive animal and mechanistic data."
There remains a need for additional research on MO A, key biomarkers that are reliable
indicators for the upstream effects elicited by the PFASs, the temporal relationship between
exposure and outcome, plus the analytical and functional impact of PFASs binding to serum
immunoglobins and/or related proteins.
8.4 Effects on Mammary Gland Development
Several studies in mice have examined postnatal mammary gland development in female
mice. A qualitative/qualitative assessment found delayed mammary gland development of
female CD-I mouse pups following maternal doses > 0.01 mg PFOA/kg in Macon et al. (2011)
and Tucker et al. (2015). Macon et al. (2011) also found significant differences from controls in
quantitative measures of longitudinal and lateral growth and numbers of terminal end buds at
1 mg/kg/day. However, Albrecht et al. (2013) found no significant differences in the average
length of mammary gland ducts and the average number of terminal end buds per mammary
gland per litter in female pups of PPARa wild type, PPARa-null, or hPPARa sv/129 following a
maternal dose of 3 mg/kg using an approach to scoring that lacked a qualitative component
adjustment such as that used by Macon et al. (2011) in identifying the 0.01 mg/kg/day dose as a
LOAEL.
The approach to scoring mammary gland development was not consistent across studies and
little information was provided on the qualitative components of the scores. This makes
comparisons across studies difficult. Statistical significance was attained at higher dose levels for
the quantitative portion of the Macon et al. (2011) scoring protocol than for the qualitative
component of the score. Tucker et al. (2015) found that CD-I mice were considerably more
sensitive to effects on mammary gland development (LOAEL 0.01 mg/kg/day) than C57BL/6
mice (NOAEL 0.1 mg/kg/day). Scoring was conducted using the Macon et al. (2011) approach.
White et al. (2011) used doses of 0 or 1 mg PFOA/kg/day for FO dams throughout gestation
with and without the addition of drinking water containing 5-ppb PFOA beginning on gestation
day 7 and continuing the contaminated drinking water during the production of two more
generations; no persistent significant differences were found in the body weights of the pups in
the Fl and F2 generations for the pups receiving 1 mg/kg/day, indicating a poor correlation
between mammary duct branching patterns and the ability to support pup growth during
lactation. The 5-mg/kg/day dose did affect body weight. Albrecht et al. (2013) also found no
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significant impacts on pup body weight in their one-generation assay at a dose of 3 mg/kg/day.
Despite the diminished ductal network assessed in the qualitative mammary gland developmental
score of the dams in White et al. (2011), milk production was sufficient to nourish growth in the
exposed pups as reflected in the body weight measurements compared to controls at the
1-mg/kg/day dose. The MOA for PFOA-induced delayed mammary gland development is
unknown and requires further investigation.
8.5 Alternative Exposure Scenarios
EPA is issuing a lifetime HA for PFOA of 0.07 |ig/L to prevent a variety of adverse
developmental effects to fetuses during pregnancy and to infants during breast feeding. Due to
the potential increased susceptibility during this critical time period, EPA used drinking water
intake and body weight parameters for lactating women to calculate the lifetime HA (see section
6.2). Specifically, EPA used the rate of 54 mL/kg-day representing the consumers only estimate
of combined direct and indirect community water ingestion at the 90th percentile for lactating
women (see Table 3-81 in USEPA [201 lb]).
As a comparative analysis, EPA calculated a lifetime HA value for alternative exposure
scenarios for the general population. Calculation of a lifetime HA value for the general
population (adults ages 21 and older) is 0.1 |ig/L, assuming a drinking water rate of 2.5 L/day
and a mean body weight of 80 kg (see Tables 3-33 and 8-1 in USEPA[201 lb]).
PFOA is extremely persistent in both the human body and the environment; thus, even a
short-term exposure results in a body burden that persists for years and can increase if additional
exposure occurs later. Human studies have shown that PFOA is transferred from mother to infant
via cord blood and breast milk. The exposure scenario for the lactating woman is the most
protective given her increased water intake rate to support milk production and thus is the basis
for EPA's recommended lifetime HA for PFOA of 0.07 |ig/L. The lifetime HA for PFOA is also
protective of adverse health effects in the adult general population (e.g., testicular and kidney
cancer, liver damage, immune effects).
8.6 Relative Source Contribution Considerations
EPA used the Exposure Decision Tree methodology to derive the RSC for this HA (USEPA
2000). Findings from studies on populations in the United States, Canada, and Western Europe
support the conclusion that diet is the major contributor to total PFOA exposure, typically with
drinking water and/or dust as important additional exposure routes, especially for sensitive
subpopulations. Estimates of relative exposure from different sources vary widely, as described
below.
• Tittlemier et al. (2007) conducted a total diet study, focused on collection and analysis of
different food items. They concluded that diet represented approximately 60% of
exposure to total PFASs, with a negligible contribution from drinking water, based on
samples collected from two cities in Canada.
• Lorber and Egeghy (2011) used models to estimate exposures for adults and 2-year-olds.
The data and analysis identify dietary ingestion as the major contributor to adult intake of
PFOA, and dust and diet for young children in different media. The authors estimated
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PFOA exposure from drinking water at 17 ng/day or approximately 24% of total intake
for both adults and children. As background concentrations of PFOA in water increase,
drinking water represents a greater source of total dietary intake.
• Gebbink et al. (2015) estimated the relative contributions of the major exposure media to
total direct and indirect PFOA exposures under assumptions of low (5th percentile),
intermediate (median values), and high (95th percentile) exposures. The authors used a
scenario-based risk assessment modeling approach with data collected in 2007 to
estimate the relative contributions of diet, dust, water, and air to total exposures. Only
data for samples collected in North America, Europe, Korea, and Japan were included in
the evaluation. The authors point out that both the blood serum concentrations and the
temporal trends of PFASs in the United States, Europe, and Japan are similar. The data
for direct and indirect contributors to serum PFOA are presented graphically in the
published paper. They are consistent with the following exposure patterns for the
combination of direct and indirect (precursor) exposures in adults:
- Low-exposure scenario: diet (-50%) > air (-25%) > dust (-15%) > water (-10%);
- Intermediate-exposure scenario: diet (-45%) > dust (-35%) > water (-10%) - air
(-10%); and
- High-exposure scenario: dust (-65%) > diet (-20%) > water (10%) > air (-5%).
As the environmental level increases, so does the contribution of precursors to total exposure,
increasing from about 15% to 30% to 60% as the exposure increases from low to high.
The approaches and assumptions used in these studies vary widely; some uncertainties
associated with these data include:
• Many of the data are obtained from review papers or individual studies conducted at
single locations often in Europe and are not nationally representative.
• Concentrations range widely in exposure estimates.
• The ambient air and dust exposure estimates are limited, regional, and variable.
• Drinking water exposure varies among age groups and individuals.
• Because of recent reductions in use of PFOA and its precursors, it is difficult to assess
current relative exposures to the general population.
Additionally, there is a lack of data on other routes of exposure:
• Estimates of dermal exposure to treated fabrics and inhalation exposure associated with
contaminated water are not available.
• Drinking water exposure estimates apply only to direct ingestion of tap water and
beverages or soups prepared locally. They do not generally include PFOA in water that
becomes incorporated in solid foods during home preparation and cooking or that is
present in commercial beverages.
• Transformation of PFOA precursors that decay or are metabolized to PFOA is a route
that is rarely evaluated in dietary studies yet can contribute to total exposure. Air and dust
can be vehicles for derivatives that metabolically degrade to PFOA.
Given these uncertainties, EPA used the Exposure Decision Tree methodology, described in
section 6.1, to estimate an RSC of 20% for drinking water for the general population.
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8.7 Sensitive Populations: Gender Differences
Some animal species have gender differences that affect toxicity of PFOA. Sexually mature
female rats excreted almost all of a 10-mg/kg dose of PFOA within 48 hours compared to only
19% excreted by male rats. Male hamsters excrete PFOA faster than female hamsters, and
female rabbits excrete PFOA slightly faster than male rabbits. Male and female mice excrete
PFOA at approximately the same rate (Hundley et al. 2006). Studies of the transporters involved
in the toxicokinetics of PFOA demonstrate that they are differentially affected by the presence of
male and female sex hormones (Cheng et al. 2006; Kudo et al. 2002). As studied in rats (Kudo et
al. 2002), the male sex hormones increased half-life (decreased excretion) of PFOA while the
female hormones were associated with shorter half-lives (increased excretion). The gender
differences in toxicokinetics in mice are not as pronounced as those in rats. Work by Cheng et al.
(2006) and Cheng and Klaassen (2009) demonstrated that the hormones affected transporters in
the liver and kidney, protecting the females and increasing the sensitivity for males. Results of
the NHANES data on PFOA suggest that in humans, serum levels are lower in females (Calafat
et al. 2007a, 2007b; Jain 2014); both menstruation and lactation are excretory routes in females
and shorten the half-life of PFOA during associated life phases.
In studies where both male and female rats were used, the males were more sensitive to
toxicity than the females (Butenhoff et al. 2004a). Mice displayed similar sensitivities following
PFOA exposure (Kennedy 1987). In the monkey studies, the number of animals per gender per
dose group was too small to reveal a difference related to gender.
Unfortunately, much work remains to be done to determine whether the gender difference
seen in rats is relevant to humans. Similarities are possible because the long half-life in humans
suggests that they might be more like the male rat than the female rat. The broad range of half-
lives in human epidemiology studies suggests a variability in human transport capabilities
resulting from the isomeric composition of the PFOA and genetic variations in transporter
structures and consequently in function (Y. Zhang et al. 2013, 2014). Genetic variation in human
transporters are identified in a review by Zai'r et al. (2008).
8.8 Sensitive Populations: Developmental Effects
PFOA-exposure during development in rats and mice resulted in increased resorptions
(mouse), increased fetal skeletal variation (rats, mouse), decreased neonatal survival (rat,
mouse), decreased postnatal body weight (mouse), delayed eye opening and body hair growth
(rat, mouse), delayed vaginal opening (mouse), accelerated preputial separation (mouse), and
delayed mammary gland development (mouse) (Butenhoff et al. 2004a; Lau et al. 2006; Macon
et al. 2011; Tucker et al. 2015; White et al. 2007, 2009, 2011; Wolf et al. 2007). Some effects
were seen as low-dose exposures such as the ossification delays and accelerated puberty in male
mice exposed via their dams to a dose of 1 mg/kg/day during gestation (Lau et al. 2006), the
mammary gland effects (0.01 mg/kg/day) (Macon et al. 2011), and the postnatal effects on body
weight in pups exposed to PFOA during gestation and lactation to doses of 3 or 5 mg/kg/day
(White et al. 2009; Wolf et al. 2007). Only the low birth weight receives support from the
epidemiology studies. The other effects generally lack correlates among the effects evaluated by
the studies.
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In the Wolf et al. (2007) study, pup postnatal body weights were lower than controls for all
exposure durations during the last 10 days of gestation evaluated. The authors found that the
magnitude of the body weight effect was directly related to the days of exposure (i.e., 3, 5, 7, or
10); the longer the exposure, the greater the body weight deficit in the male and female pups
during the PND 2-22 time period. In male but not female pups, the exposure duration deficits in
body weight persisted up to PND 92. The difference in the male rat response over the PND 29-
92 period likely reflects their longer half-life than females.
Both gestational and lactational exposures contribute to the impact of PFOA on body weight
during early life as illustrated by cross-fostering control unexposed female pups with those dosed
with PFOA. Three cross-fostering combinations were evaluated by White et al (2009): control
pups nursed by exposed dams, exposed pups nursed by control dams, and exposed pups nursed
by exposed dams. Two doses were evaluated: 3 and 5 mg/kg/day. The PND 1-10 body weight
data were only provided for the 5-mg/kg/day dose. PFOA exposures significantly reduced pup
body weights and increased liver weights. The body weight deficits compared to control were
greatest for the gestation and lactation exposure combination and lowest for the lactation-only
group.
Diet can influence the risk associated with PFOA exposures. Animal studies demonstrate an
increased risk for liver steatosis in animals on a high-fat diet and possibly for insulin resistance
(Hines et al. 2009; Quist et al. 2015; Tan et al. 2013). The epidemiology data are not supportive
of a correlation with insulin resistance, but the observations of elevated serum triglycerides,
especially among a highly exposed population, could be viewed as a risk factor for steatosis.
Most of the epidemiology studies did not evaluate dietary factors as part of the study design for
either birth weight or serum lipids (e.g., cholesterol, triglycerides, LDL).
9.0 ANALYTICAL METHODS
EPA developed a liquid chromatography/tandem mass spectrometry (LC/MS/MS) analytical
method— Method 537—to monitor drinking water for 14 select perfluorinated alkyl acids that
include PFOA (USEPA 2009b). Accuracy and precision data were generated for PFOA, PFOS,
and the other 12 PFASs in reagent water, finished ground water, and finished surface water. This
method identifies a single laboratory lowest concentration minimum reporting level or
quantitation limit for PFOA at 5.1 ng/L (0.0051 |ig/L) and for PFOS at 6.5 ng/L (0.0065 |ig/L).
The method-published detection limit for PFOA is 1.7 ng/L (0.0017 |ig/L).
In this method, PFAS standards, extracts, and samples should not come into contact with any
glass containers or pipettes because PFASs can potentially adsorb to the surface of the glassware.
Polypropylene containers should be used instead. Also, these compounds can be found in
commonly used laboratory supplies and equipment, such as PTFE products, liquid
chromatograph solvent lines, methanol, aluminum foil, and solid phase extraction (SPE) sample
transfer lines. These materials need to be routinely demonstrated to be free of interferences per
the guidelines for laboratory reagent blanks described in the method. In summary, the method
procedure involves passing a preserved 250-mL water sample (fortified with an extraction
surrogate) through a SPE cartridge containing polystyrenedivinylbenzene (SDVB) to extract the
method analytes and surrogates.
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The compounds are eluted from the SPE with a small amount of methanol. The extract is
concentrated to dryness with nitrogen in a heated water bath, and then adjusted to a 1-mL
volume with 96%:4% (vol/vol) methanol:water after adding the internal standards. The extract is
injected into a liquid chromatograph that is interfaced to an MS/MS. The analytes are separated
and identified by comparing the acquired mass spectra and retention times to reference spectra
and retention times for calibration standards acquired under identical LC/MS/MS conditions. The
concentration of each analyte is determined by using the internal standard technique. Surrogate
analytes are added to all field and quality control samples to monitor the extraction efficiency of
the method analytes. Method 537: Determination of Selected PerfluorinatedAlkyl Acids in
Drinking Water by Solid Phase Extraction and Liquid Chromatography/Tandem Mass
Spectrometry (LC/MS/MS) (USEPA 2009b) is available for download at
http://www.epa.gov/nerlcwww/ordmeth.htm.
10.0 TREATMENT TECHNOLOGIES
As mentioned above, PFOA is an organic compound in which the carbon-hydrogen bonds are
replaced by carbon-fluorine bonds. This influences the chemical characteristics of both
molecules and, therefore, will impact the effectiveness of any given drinking water treatment
process. The characteristics of organic contaminants that treatment processes take advantage of
include molecular size, solubility, ionic form, volatility, oxidizability, hydrolysis, photolysis, and
biodegradability. Because fluorine is the most electronegative element, the carbon-fluorine bond
will be one of the strongest bonds in nature, making it exceedingly resistant to biodegradation,
hydrolysis, oxidation, and photolysis. Also, because PFOA is a dissolved contaminant that is
resistant to oxidation to an insoluble form, treatment processes that are designed for paniculate
control such as conventional treatment will not be effective. This leaves adsorption, ion
exchange resins, and high-pressure membranes as the technologies that can be effective. The
following subsections discuss the effectiveness of commonly used drinking water technologies in
rough order of applicability for PFOA and PFOS removal. Additional information can be found
on EPA's Drinking Water Treatability Database (USEPA 2015b) at
https://iaspub.epa.gov/tdb/pages/general/home.do.
To varying degrees of applicability, the technologies discussed below can be employed in
centralized drinking water facilities or in a distributed fashion such as point-of-entry (POE) or
point-of-use (POU) applications in buildings and homes. As they imply, POE systems treat the
water as it enters the building or house, and POU systems treat the water where used, such as a
kitchen or bathroom sink. Although the cost of treatment varies with scale, the following general
discussion on the relative effectiveness of each technology applies regardless of scale. One
reference below specifically addresses POU systems (MDH 2008).
Activated Carbon Adsorption
Activated carbon is applied in either powdered or granular form. Either can be effective;
however, because PFOA and PFOS have moderate adsorbability, the specifics of the design are
very important for achieving successful treatment.
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Powdered Activated Carbon
Powdered activated carbon (PAC) is often applied prior to, or within a, conventional
treatment train. The contaminant-loaded PAC is then removed along with the other particulates.
Although some studies have shown limited PFOA and PFOS removal in plants using PAC
(Quifiones and Snyder 2009), in general, PAC can be an effective treatment strategy for the
removal of PFOA and PFOS given the correct choice of carbon type, high enough carbon doses,
and adequate contact time (Dudley et al. 2015; Hansen et al. 2010).
Granular Activated Carbon
Granular activated carbon is applied as a filtration step either as a filter adsorber where a
relatively short carbon cap is added to an existing sand filter, or as a post-filter adsorber where a
deeper bed is employed as a stand-alone unit following a typical sand filter. Because PFOA and
PFOS have moderate adsorbability, a post-filter adsorber with a deeper bed is considered a safer
approach. In general, granular activated carbon treatment was found to be effective given the
correct choice of carbon, adequate bed depth, moderate or low hydraulic loading rate, and
frequent replacement or regeneration of the carbon (Appleman et al. 2013, 2014; MDH 2008;
Shivakoti et al. 2010; Takagi et al. 2008).
Membrane Technologies
There are many types of membrane technologies. They can be broadly classified as either
low-pressure or high-pressure systems. This distinction also corresponds to the general
effectiveness of removing PFOA and PFOS with low-pressure membranes being ineffective,
while high-pressure membranes are effective.
Low-pressure Membranes
Low-pressure systems incorporating cartridge, microfiltration, or ultrafiltration membranes
are designed for paniculate control. They have relatively large pore structures through which
water and dissolved contaminants can easily flow, leaving behind larger particulate matter that
includes turbidity and microbiological agents. Low-pressure membranes have been found to be
ineffective for PFOA and PFOS control (McLaughlin et al. 2011; Thompson et al. 2011). This is
consistent with other treatment processes (e.g., conventional treatment) that target particulate
contaminants but not dissolved contaminants. However, as with conventional treatment, low-
pressure membranes can be effective if they are used in conjunction with PAC. The PAC will
adsorb the PFOA and PFOS, and the low-pressure membrane will remove the spent PAC. Care
should be taken in the design of the system, including the choice of the PAC as mentioned above
(Dudley etal. 2015).
High-pressure Membranes
High-pressure systems have a much tighter pore structure, relying on water diffusing through
the membrane material. High-pressure systems such as nanofiltration and reverse osmosis can
reject not only particulates, but also dissolved constituents such as organic contaminants and
salts. Reverse-osmosis membranes are the tightest of the high-pressure systems, having the
ability to reject monovalent salts such as sodium chloride (e.g., sea water desalination). High-
pressure membrane systems have been shown to be very effective for PFOA and PFOS
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) - May 2016 66
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(Appleman et al. 2013, 2014; MDH 2008; Quifiones and Snyder 2009; Tang et al. 2006, 2007;
Thompson et al. 2011).
Ion Exchange Resin Treatment
There are two broad categories of ion exchange resins: cationic and anionic. Cationic
exchange resins are effective for removing positively charged contaminants. Anion exchange
resins are effective for negatively charged contaminants. Because PFOA and PFOS are
negatively charged in drinking waters, cation-exchange resins will not be effective, and
therefore, have not been studied. There have been studies that have evaluated different anion
exchange resins (macroporous styrenedivinylbenzene, gel-type polystyrene divinylbenzene, and
polyacrylic quaternary amine resins). Generally, anion exchange resins have been found to be
effective for PFOA and PFOS removal (Appleman et al. 2014; Carter and Farrell 2010;
Chularueangaksorn et al. 2013; Dudley et al. 2015), although the design of the system including
regeneration effectiveness is important. Special consideration should be given to dealing with the
regenerate brine waste, and if frequent regenerations are needed, to the amount of operator effort
and expertise required.
Oxidation /Disinfection
Oxidation/disinfection processes can transform certain contaminants into different molecules,
which ideally have less toxicity. It also can transform certain dissolved constituents into a higher
oxidation state that might be less soluble (e.g., iron, manganese). The less soluble form can then
be precipitated and removed in the floe or on a media filter of a conventional treatment system.
Due to the strength of the carbon-fluorine bond, all drinking water oxidants or disinfectants have
been shown to be ineffective in reacting PFOA or PFOS. This has been shown numerous times
for common oxidative/disinfection agents such as packed tower aeration, chloramination,
chlorination, ozonation, potassium permanganate, and ultraviolet (UV) treatment (Appleman et
al. 2014; Hori et al. 2004; Liu et al. 2012; McLaughlin et al. 2011; Quifiones and Snyder 2009;
Schroder and Meesters 2005; Shivakoti et al. 2010; Thompson et al. 2011). It also is true for
advanced oxidation processes (AOPs) that use the nonselective hydroxyl radical as an oxidative
agent. There are many ways of producing hydroxyl radicals, usually combining technologies
such as hydrogen peroxide plus iron (Fenton's reagent), ozone plus peroxide, UV plus titanium
dioxide, UV plus ozone, and UV plus peroxide. All of these combinations have been shown to be
ineffective for PFOA and PFOS control at reasonable contact times (Benotti et al. 2009; Hori et
al. 2004; Schroder and Meesters 2005; Tellez 2014).
Biological Treatment
Similar to the discussion on oxidation processes, because of the strength of the carbon-
fluorine bond, it is expected that both aerobic and anaerobic biological treatment processes
(e.g., biofiltration, bioreactors) are expected to be ineffective for PFOA and PFOS removal. A
number of researchers have found this to be the case (Kwon et al. 2014; Saez et al. 2008;
Thompson et al. 2011). Some results have shown that specific microbes could have the ability to
break the carbon-to-carbon bonds in PFOS, albeit slowly; however, this cannot be engineered
into a consistent and robust treatment process (Kwon et al. 2014).
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Conventional Treatment
Conventional treatment is commonly defined as a series of successive steps: rapid mix,
coagulation, flocculation, sedimentation, and filtration. Certain variations exist, such as direct
filtration, which does not employ a sedimentation step. Regardless of the configuration,
conventional treatment is designed to remove particulates (e.g., turbidity and microbiological
agents). Dissolved contaminants, however, will not be removed by conventional treatment. The
exception is when the contaminants are first oxidized to an insoluble form (e.g., iron,
manganese), or if they are exceedingly hydrophobic as evidenced by an extremely low solubility.
Therefore, because of the resistance of PFOA and PFOS to oxidation to an insoluble form and
their moderately high solubility, conventional treatment is not expected to be effective in their
removal, even in enhanced coagulation conditions. Numerous studies have confirmed this
statement (Appleman et al. 2014; Loos et al. 2007; Quifiones and Snyder 2009; Shivakoti et al.
2010; Skutlarek et al. 2006; Tabe et al. 2010; Takagi et al. 2008; Thompson et al. 2011; Xiao et
al. 2013).
Similar to low-pressure membranes, conventional treatment can be effective if it is used in
conjunction with PAC (see above). The PAC will adsorb the PFOA and PFOS, and the
conventional treatment system will remove the spent PAC in the sedimentation and filtration
steps. Care should be taken in the design of the system, including the choice of the PAC (Dudley
etal. 2015).
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12.0 APPENDIX A-QUANTITATIVE CANCER ASSESSMENT MODELING
Multistage Model for Leydig Cell Tumors
Multistage Cancer Model. (Version: 1.9; Date: 05/26/2010)
Input Data File: C:/IData/MyFiles/PFOA-PFOS/PFOA Docs/msc_Leydig_Opt.(d)
Gnuplot Plotting File: C:/IData/MyFiles/PFOA-PFOS/PFOA Docs/msc_Leydig_Opt.pit
Thu May 09 11:59:27 2013
BMDS_Model_Run
The form of the probability function is:
P [response] = background + (1-background) * [1-EXP (-betal*dose/xl-beta2*dose/x2) ]
The parameter betas are restricted to be positive
Dependent variable = Col2
Independent variable = Coll
Total number of observations = 3
Total number of records with missing values = 0
Total number of parameters in model = 3
Total number of specified parameters = 0
Degree of polynomial = 2
Maximum number of iterations = 250
Relative Function Convergence has been set to: le-008
Parameter Convergence has been set to: le-008
Default Initial Parameter Values
Background = 0.0132945
Beta(l) = 0.0097738
Beta(2) = 0
Asymptotic Correlation Matrix of Parameter Estimates
( *** The model parameter(s) -Beta(2)
have been estimated at a boundary point, or have been specified by
the user,
and do not appear in the correlation matrix )
Background Beta(l)
Background 1 -0.64
Beta(l) -0.64 1
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) - May 2016 99
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Parameter Estimates
Variable
Background
Beta(l)
Beta(2)
Estimate
0.00409839
0.0116288
0
Std. Err.
*
*
*
95.0% Wald Confidence Interval
Lower Conf. Limit Upper Conf. Limit
Indicates that this value is not calculated.
Model
Full model
Fitted model
Reduced model
AIC: 62.6936
Log(likelihood)
-28.6454
-29.3468
-34.0451
Analysis of Deviance Table
Deviance
# Param's
3
2
1
1.40286
10.7995
Test d.f.
1
2
P-value
0.2362
0.004518
Goodness of Fit
Dose
0.0000
1.3000
14.2000
Est. Prob.
0.0041
0.0190
0.1557
Expected
0.205
0.952
7.784
Observed
0.000
2.000
7.000
Size
50
50
50
Scaled
Residual
-0.454
1.084
-0.306
ChiA2 = 1.48 d.f. = 1
P-value = 0.2245
Benchmark Dose Computation
Specified effect = 0.04
Risk Type = Extra risk
Confidence level = 0.95
BMD = 3.51044
BMDL = 1.99346
BMDU = 10.7788
Taken together, (1.99346, 10.7788) is a 90% two-sided confidence interval for the BMD
Multistage Cancer Slope Factor = 0.0200656
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) - May 2016
100
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Multistage Cancer Model with 0.95 Confidence Level
I
C
o
0.3
0.25
0.2
0.15
0.1
0.05
Multistage Cancer
Linear extrapolation
BMDL
BMD
6 8
dose
10
12
14
11:5905/092013
Multistage Cancer Model. (Version: 1.9; Date: 05/26/2010)
Input Data File: C:/IData/MyFiles/PFOA-PFOS/PFOA Docs/msc_Leydig_Opt.(d)
Gnuplot Plotting File: C:/IData/MyFiles/PFOA-PFOS/PFOA Docs/msc_Leydig_Opt.pit
Thu May 09 12:05:42 2013
BMDS_Model_Run
The form of the probability function is:
P[response] = background + (1-background)*[1-EXP(-betal*dose/xl)]
The parameter betas are restricted to be positive
Dependent variable = Col2
Independent variable = Coll
Total number of observations = 3
Total number of records with missing values = 0
Total number of parameters in model = 2
Total number of specified parameters = 0
Degree of polynomial = 1
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) - May 2016
101
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Maximum number of iterations = 250
Relative Function Convergence has been set to: le-008
Parameter Convergence has been set to: le-008
Default Initial Parameter Values
Background =
Beta(l) =
0.0132945
0.0097738
Asymptotic Correlation Matrix of Parameter Estimates
Background
Beta(l)
Background
1
-0.64
Beta (1)
-0.64
1
Variable
Background
Beta(1)
Parameter Estimates
Estimate
0.00409839
0.0116288
Std. Err.
95.0% Wald Confidence Interval
Lower Conf. Limit Upper Conf. Limit
Indicates that this value is not calculated.
Analysis of Deviance Table
Model Log(likelihood) # Param's Deviance Test d.f. P-value
Full model -28.6454 3
Fitted model -29.3468 2 1.40286 1 0.2362
Reduced model -34.0451 1 10.7995 2 0.004518
AIC: 62.6936
Dose
Est. Prob.
Goodness of Fit
Expected Observed
Size
=1.48
d.f. = 1
P-value = 0.2245
Scaled Residual
0.0000
1.3000
14.2000
0.0041
0.0190
0.1557
0.205
0.952
7.784
0.000
2.000
7.000
50
50
50
-0.454
1.084
-0.306
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) - May 2016
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Benchmark Dose Computation
Specified effect = 0.04
Risk Type = Extra risk
Confidence level = 0.95
BMD = 3.51044
BMDL = 1.99346
BMDU = 8.7003
Taken together, (1.99346, 8.7003) is a 90% two-sided confidence interval for the BMD
Multistage Cancer Slope Factor = 0.0200657
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) - May 2016 103
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