Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 1 

Drinking Water Health 

Advisory for 

Perfluorooctanoic Acid 

(PFOA) 

United States 

Environmental Protection 

Agency 

Office of Water

Mail Code 4304T

EPA 822-R-16-005

May 2016
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 2 

Drinking Water Health Advisory

for Perfluorooctanoic Acid (PFOA)

Prepared by: 

U.S. Environmental Protection Agency 

Office of Water (4304T)

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 3 

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 4 

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 Uses of PFOA .................................................................................................................. 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 for PFOA ......................................... 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
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 5 

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
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 6 

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
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 7 

ABBREVIATIONS AND ACRONYMS

ALT alanine aminotransferase

ALP alkaline phosphatase

APFO ammonium perfluorooctanoate

AST aspartate aminotransferase

AUC area under the curve 

BAF bioaccumulation factor 

BCF bioconcentration factor 

BMDL benchmark dose level

BMF biomagnification factor 

bw body weight 

CAR constitutive androstane receptor

CCK cholecystokinin 

CCL Contaminant Candidate List

COPD chronic obstructive airways disease

CWA Clean Water Act

DWEL drinking water equivalent level

DWI drinking water intake

EPA U.S. Environmental Protection Agency 

FXR farnesoid X receptor

GFR glomerular filtration rate

GGT gamma-glutamyl transferase

HA Health Advisory 

HDL high-density lipoprotein

HED human equivalent dose 

HESD Health Effects Support Document 

IgM immunoglobulin M 

IRIS Integrated Risk Information System 

Koc organic carbon-water partitioning coefficient

Kow octanol-water partition coefficient

LCT Leydig cell tumor 

LC/MS/MS liquid chromatography/tandem mass spectrometry

LDL low-density lipoprotein

LOAEL lowest observed adverse effect level

MOA mode of action 

MRL minimum reporting level

ng/L nanograms per liter

NHANES National Health and Nutrition Examination Survey 

NOAEL no observed adverse effect level

PAC powdered activated carbon 

PACT pancreatic acinar cell tumors

PBPK physiologically based pharmacokinetic model 

PFAS perfluoroalkyl substance 

PFC perfluoroinated compounds 

PFOA perfluorooctanoic acid

PFOS perfluorooctanesulfonic acid
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 8 

PTFE polytetrafluoroethylene

pg/L picograms per liter

PND post-natal day

POD point of departure 

POE point-of-entry

POU point-of-use 

PPARα peroxisome proliferator activated receptor alpha 

PWS public water system

PXR pregnane X receptor

REACH Registration, Evaluation, Authorization, and Restriction of Chemicals 

RfD reference dose

RSC relative source contribution

SDWA Safe Drinking Water Act

SNUR Significant New Use Rule

SRBC sheep red blood cell 

TMF trophic magnification factor 

TNSSS Total National Sewage Sludge Survey 

UCMR 3 third Unregulated Contaminant Monitoring Rule

UF uncertainty factor 

UV ultraviolet 
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 9 

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 (μg/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 10

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% of the RfD.

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 μg/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 μg/L). 

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. 
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 11

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.1

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 12

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 [μg/L]) by 

approximately 2% of PWSs nationwide. To-date, PFOA has been measured above the new 

lifetime HA level of 0.07 μg/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 μg/L. For the 

latest UCMR 3 results, please refer to https://www.epa.gov/dwucmr/occurrence-dataunregulated-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 (μg/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 13

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

Guideline Value

(μg/ L) Source

Delaware Department of Resources and Environmental Control 0.4 DNREC (2016)

Maine Department of Health and Human Services 0.1 Maine DHHS (2014)

Michigan Department of Environmental Quality 0.42 Michigan DEQ (2013)

Minnesota Department of Health 0.3 MDH (2009)

New Jersey Department of Environmental Protection 0.04 NJDEP (2014)

North Carolina Division of Water Quality 2 NCDEQ (2013)

Vermont Agency of Natural Resources 0.02 Vermont ANR (2016)

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

Guideline Value (μg/ L)

Health-based Administrative Source

German Ministry of 

Health 0.3 Composite precautionary guidance value for PFOA+PFOS is 0.1 

German Ministry of 

Health (2006)

United Kingdom (UK) 

Drinking Water 

Inspectorate

5.0

Action levels:

Tier 1: potential hazard

Tier 2: > 0.3

Tier 3: > 5.0

Tier 4: > 45

UK Drinking Water 

Inspectorate (2009)

Danish Ministry of the 

Environment 0.3 

Composite drinking water criteria are based 

on relative toxicity of PFOS, PFOA, and 

PFOSA

Danish Ministry of the 

Environment (2015)
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 14

Country/Agency

Guideline Value (μg/ L)

Health-based Administrative Source

Swedish National Food 

Agency --

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

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 perfluoroalkylbased 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-undertsca/long-chain-perfluorinated-chemicals-pfcs. 
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 15

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., TeflonTM). 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

 

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 16

Table 2-1. Chemical and Physical Properties of PFOA

Property Perfluorooctanoic Acid Source

Chemical Abstracts Service 

Registry No. (CASRNa) 

335-67-1

Chemical Abstracts Index 

Name

2,2,3,3,4,4,5,5,6,6,7,7,8,8,8-

pentadecafluorooctanoic acid

Synonyms PFOA; Pentadecafluoro-1-octanoic acid; 

Pentadecafluoro-n-octanoic acid; 

Octanoic acid, pentadecafluoro-; 

Perfluorocaprylic acid; 

Pentadecafluorooctanoic acid; 

Perfluoroheptanecarboxylic acid; 

Chemical Formula C8HF15O2

Molecular Weight (g/mol) 414.09 HSDB (2012); Lide (2007); SRC (2016)

Color/Physical State White powder (ammonia salt) HSDB (2012); Lewis (2004)

Boiling Point 192.4oC; Stable when bound HSDB (2012); Lide (2007); SRC (2016)

Melting Point 54.3 oC HSDB (2012); Lide (2007); SRC (2016)

Vapor Pressure 0.525 mm Hg at 25 °C (measured)

0.962 mm Hg at 59.25 °C (measured)

Hekster et al. (2003); HSDB (2012); 

SRC (2016)

ATSDR (2015); Kaiser et al. (2005)

Henry’s Law Constant Not measureable ATSDR (2015)

pKa 2.80 SRC (2016)

Koc 2.06 Higgins and Luthy (2006)

Kow Not measurable ATSDR (2015); EFSA (2008)

Solubility in Water 9.50 x 103 mg/L at 25 oC (estimated) ATSDR (2015); Hekster et al. (2003);

HSDB (2012); Kauck and Diesslin 

(1951); SRC (2016)

Half-life in Water (25oC) Stable UNEP (2015)

Half-life in Air Stable when bound UNEP (2015)

Notes: 

Kow = octanol-water partition co-efficient; Koc = organic carbon-water partitioning coefficient; g/mol = grams per mole a

The 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. 2011a; 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). 
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 17

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. 2011b; 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 μg/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 μg/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. 
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 18

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 μg/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 μg/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 μg/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 μg/L. One system had an average 

concentration of 0.09 μg/L and a maximum reported concentration of 0.25 μg/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 μg/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 μg/L in the municipal water distribution system and up 

to 14 μg/L in private wells between 2002 and 2005 (Emmett et al. 2006). A study in Minnesota 

reported PFOA concentrations up to 0.9 μg/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. 
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 19

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 entrées, 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 20

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 

of fish does not reduce the levels of PFOA in the fish (or the consumer’s dietary exposure) 

(Bhavsar et al. 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 

(Blaine et al. 2014). 
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 21

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 [μg/dm2] paper) of PFOA 

into food oil before cooking and another 7 μg/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) (TeflonTM). PFOA is a processing aid in the manufacture of 

PTFE. Begley et al. (2005) also evaluated migration of PFOA to foods from cooking in 

TeflonTM-lined cookware and found it to be much lower (0.03 μg/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 22

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 23

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 2011a). 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. 2011b; USEPA 2011a; Washington et al.

2010a, 2010b). 

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 μg/kg dry 

weight), followed by PFOA (mean 34 ± 22 μg/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 μg/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 μg/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. 
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 24

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., StainmasterTM, ZonylTM, NuvaTM, UnidyneTM, BaygardTM) 

(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., TeflonTM) 

• 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-TexTM) 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).
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 25

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-
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 26

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 [PPARα], 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-1 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). 
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 27

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 μg/L, and the 95th percentile concentration decreased from 11.9 to 5.7 μg/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; Völkel 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). 
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 28

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). 

Figure 3-1. Conceptual Model for PFOA in Finished Drinking Water
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 29

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 particulate 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 
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 30

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 PPARα receptors, 

making it difficult to determine if this change is a reflection of PPARα activation or an indication 

of PFOA toxicity. The PPARα 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.
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 31

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.
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 32

• 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) A 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 threestep 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 
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 33

concentration representing either an NOAEL or LOAEL experimental dose after applying 

uncertainty factors established following EPA guidelines. 

RfD = HED NOAEL or HED 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. 

DWEL = RfD x bw

DWI

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 11 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.
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 34

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). 
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 35

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)10 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 PFOAexposed males also had increased activity counts compared to control males in circadian activity 
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 36

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 PPARα 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 PPARα. 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 PPARα-related response. 

DeWitt et al. (2008) used different functionality assays in their study in C57Bl/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, C57Bl/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. 
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 37

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 PPARα 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-1 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 F1 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 SpragueDawley 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, F0 and F1 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-toterminal 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-1 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.
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 38

Conversely, Quist et al. (2015) found no dose-related impact on serum leptin in CD-1 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 PPARα activation in rodents. PFOA is an activator of the PPARα 

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 PPARα 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 PPARα 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). 
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 39

Cholesterol and/or triglycerides were monitored in few animal studies. Nakamura et al. 

(2009) found that mice with a normal PPARα 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 PFOAtreated mice with a humanized PPARα receptor or PPARα 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 PPARα 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. 
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 40

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) (FitzSimon 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 [μg/mL]), and the mean serum levels in the C8 

population studies were around 0.08 μg/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 μg/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 ln 

ALT and ln 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 ln PFOA concentration with increasing serum 

ALT and ln 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 
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 41

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 μg/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 μg/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 
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 42

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 μg/mL) in women 

and decreases in fecundity and fertility (Fei et al. 2009; Vélez et al. 2015), but there are no clear 

effects of PFOA on male fertility endpoints (0.0035–0.005 μg/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. 
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 43

• 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; Wu et al. 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 halflife

• 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 PPARα 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 44

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 PPARα 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-1 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-1 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), Leydig 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 
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 45

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 μg/mL) and kidney and testicular cancers (Barry et al. 2013; Vieira et al. 2013); 
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 46

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 μg/mL and colorectal, breast, prostate, bladder, 

and liver cancer (Bonefeld-Jørgensen et al. 2014; Eriksen et al. 2009; Hardell et al. 2014; Innes 

et al. 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 PPARα 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 

PPARα-null mice dosed with PFOA for up to 7 days. This study showed that animal responses 

were consistent with PPARα agonism, but evidence also shows PPARγ agonism (downregulation 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 PPARα, few studies have 

evaluated whether additional steps (i.e., cell proliferation and apoptosis) are in the hypothesized 

PPARα 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 PPARα-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 MOAs 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 wildtype, PPARα- null and mice with humanized PPARα. 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 47

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 PPARα, γ, δ 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 Leydig 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 2B) 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. 
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 48

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 HED 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 HED, 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 10–4 L/kg bw/day by the following equation: 

CL = Vd x (ln 2 ÷ t1⁄2) = 0.17 L/kg bw x (0.693 ÷ 839.5 days) = 0.00014 L/kg bw/day 

Where:

Vd = 0.17 L/kg

ln 2 = 0.693 

t1⁄2 = 839.5 days (2.3 years x 365 days/year = 839.5 days) 
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 49

Multiplying the derived average serum concentrations (in μg/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 Dosing 

duration

days

NOAEL

mg/kg/d 

NOAEL

Av serum

mg/L

HED

mg/kg/d

LOAEL

mg/kg/d

LOAEL

(Av serum)

mg/L

HED

mg/kg/d

DeWitt et al. (2008): 

mice; ↓ IgM response 

to SRBC

15 1.88 38.2 0.0053 3.75 61.9 0.0087

Lau et al. (2006): mice

decreased ↓ pup 

ossification (m, f), 

accelerated male 

puberty

17 None - - 1 38.0 0.0053

Palazzolo et al. (1993); 

Perkins et al. (2004): 

rats; ↑liver 

weight/necrosis

91 0.64 31.6 0.0044 1.94 77.4 0.0108

Wolf et al. (2007): 

mice; GD 1–17

↓Pup body weight

17 None - - 3 77.9 0.0109

Wolf et al. (2007): 

mice; GD 7–17

↓Pup body weight1

11 None - - 5 87.9 0.0123

Butenhoff et al.

(2004a): ↓ relative 

body weight/↑ relative 

kidney weight and 

↑kidney: brain weight 

ratio in F0 and F1 at 

sacrifice

84 None - - 1 45.9 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 μg/mL (mouse) to 

87.6 μg/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). 
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 50

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 F0 and F1 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 (UFH) 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 HED was not adjusted 
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 51

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 (UFD) 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 twogeneration 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 UFH, 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 52

Table 5-2. Candidate RfDs Derived from the HEDs from the Pharmacokinetic Model 

Average Serum Values

POD 

HED POD

mg/kg/day UFH UFA UFL UFS UFD UFtotal

Candidate RfD

mg/kg/day

PK-HEDNOAEL Palazzolo

et al. (1993)/Perkins et al. 

(2004)

rats; ↑liver 

weight/necrosis

0.0044 10 3 - - - 30 0.00015

PK-HEDLOAEL Wolf et al. 

(2007) GD1-17 mice; 

↓Pup body weight

0.0109 10 3 10 - - 300 0.00004

PK-HEDLOAEL Wolf et al. 

(2007) GD 7-17 mice; 

↓Pup body weight (serum 

from pups on PND 22)

0.0123 10 3 10 - - 300 0.00004

PK-HEDNOAEL DeWitt et 

al. (2008)

mice; ↓ IgM response to 

SRBC

0.0053 10 3 - 10 - 300 0.00002

PK-HEDLOAEL Lau et al. 

(2006)

mice decreased ↓ pup 

ossification (m, f), 

accelerated male puberty

0.0053 10 3 10 - - 300 0.00002

PK-HEDLOAEL Butenhoff

et al. (2004a)

↓ relative body weight/↑ 

relative kidney weight 

and ↑kidney: brain weight 

ratio in F0 and F1 at 

sacrifice

0.0064 10 3 10 - - 300 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; UFH = 

intraindividual uncertainty factor; UFA = interspecies uncertainty factor; UFS = subchronic to chronic uncertainty factor; UFL = 

LOAEL to NOAEL uncertainty factor; UFD = 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 F0 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 
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 53

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 F1 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. 
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 54

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 

F0 and F1 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 2011b). 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. 
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 55

DWEL = RfD x bw

DWI

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 

USEPA 2011b). 

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 μg/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 μg/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 
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 56

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 μg/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 wellconducted 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 (BMDL04) was 

1.99 mg/kg/day, which yields a HED 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 2011b) 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 2011b). The resultant 0.5 μg/L value is greater than the lifetime HA (0.07 μg/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 μg/L

(0.07 mg/kg/day x 2.5L/d) 
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 57

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 doseresponse 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. 
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 58

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 
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 59

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, 
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 60

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 HESDs 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 MOA, 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-1 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 PPARα wild type, PPARα-null, or hPPARα 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-1 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 F0 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 F1 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 
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 61

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 μg/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 [2011b]).

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 μg/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[2011b]). 

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 μg/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 
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 62

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. 
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 63

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 halflives 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 Zaïr 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. 
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 64

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 μg/L) and for PFOS at 6.5 ng/L (0.0065 μg/L). 

The method-published detection limit for PFOA is 1.7 ng/L (0.0017 μg/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.
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 65

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 Perfluorinated Alkyl 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 particulate 

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. 
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 66

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 

(Quiñones 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 particulate 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, lowpressure 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 et al. 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). Highpressure membrane systems have been shown to be very effective for PFOA and PFOS 
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 67

(Appleman et al. 2013, 2014; MDH 2008; Quiñones 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 floc 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; Quiñones and Snyder 2009; 

Schröder 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; Schröder and Meesters 2005; Tellez 2014). 

Biological Treatment 

Similar to the discussion on oxidation processes, because of the strength of the carbonfluorine 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). 
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 68

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; Quiñones 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 

et al. 2015). 
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 69

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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:/1Data/MyFiles/PFOA-PFOS/PFOA Docs/msc_Leydig_Opt.(d) 

 Gnuplot Plotting File: C:/1Data/MyFiles/PFOA-PFOS/PFOA Docs/msc_Leydig_Opt.plt 

 Thu May 09 11:59:27 2013 

 ==================================================================== 

 BMDS_Model_Run 

~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~ 

 The form of the probability function is: 

 P[response] = background + (1-background)*[1-EXP(-beta1*dose^1-beta2*dose^2)] 

 The parameter betas are restricted to be positive 

 Dependent variable = Col2 

 Independent variable = Col1 

 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: 1e-008 

 Parameter Convergence has been set to: 1e-008 

Default Initial Parameter Values 

Background = 0.0132945

Beta(1) = 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(1)

Background 1 -0.64

Beta(1) -0.64 1
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 100

Parameter Estimates 

Variable Estimate Std. Err.

95.0% Wald Confidence Interval

Lower Conf. Limit Upper Conf. Limit

Background 0.00409839 * * *

Beta(1) 0.0116288 * * *

Beta(2) 0 * * *

* - 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 

Goodness of Fit 

Dose Est._Prob. Expected Observed Size

Scaled 

Residual

0.0000 0.0041 0.205 0.000 50 -0.454

1.3000 0.0190 0.952 2.000 50 1.084

14.2000 0.1557 7.784 7.000 50 -0.306

Chi^2 = 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 101

 ==================================================================== 

 Multistage Cancer Model. (Version: 1.9; Date: 05/26/2010) 

 Input Data File: C:/1Data/MyFiles/PFOA-PFOS/PFOA Docs/msc_Leydig_Opt.(d) 

 Gnuplot Plotting File: C:/1Data/MyFiles/PFOA-PFOS/PFOA Docs/msc_Leydig_Opt.plt 

 Thu May 09 12:05:42 2013 

 ==================================================================== 

 BMDS_Model_Run 

~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~ 

 The form of the probability function is: 

 P[response] = background + (1-background)*[1-EXP(-beta1*dose^1)] 

 The parameter betas are restricted to be positive 

 Dependent variable = Col2 

 Independent variable = Col1 

 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 

 0

 0.05

 0.1

 0.15

 0.2

 0.25

 0.3

 0 2 4 6 8 10 12 14

Fraction Affected

dose

Multistage Cancer Model with 0.95 Confidence Level

11:59 05/09 2013

BMDL BMD

 

Multistage Cancer

Linear extrapolation
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 102

 Maximum number of iterations = 250 

 Relative Function Convergence has been set to: 1e-008 

 Parameter Convergence has been set to: 1e-008 

Default Initial Parameter Values 

Background = 0.0132945

Beta(1) = 0.0097738

Asymptotic Correlation Matrix of Parameter Estimates 

Background Beta(1)

Background 1 -0.64

Beta(1) -0.64 1

Parameter Estimates 

Variable Estimate Std. Err.

95.0% Wald Confidence Interval

Lower Conf. Limit Upper Conf. Limit

Background 0.00409839 * * *

Beta(1) 0.0116288 * * *

* - 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 

Goodness of Fit 

Dose Est._Prob. Expected Observed Size Scaled Residual

0.0000 0.0041 0.205 0.000 50 -0.454

1.3000 0.0190 0.952 2.000 50 1.084

14.2000 0.1557 7.784 7.000 50 -0.306

Chi^2 = 1.48 d.f. = 1 P-value = 0.2245 
Drinking Water Health Advisory for Perfluorooctanoic Acid (PFOA) – May 2016 103

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