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DRAFT RISK ASSESSMENT 

OF THE POTENTIAL HUMAN HEALTH EFFECTS 

ASSOCIATED WITH EXPOSURE TO

PERFLUOROOCTANOIC ACID

AND ITS SALTS

U.S. Environmental Protection Agency

Office of Pollution Prevention and Toxics

Risk Assessment Division

January 4, 2005
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Table of Contents

Executive Summary 6 

1.0 Scope of the Assessment 11

2.0 Chemical Identity 12

2.1 Physicochemical Properties 12

3.0 Hazard Characterization 13

3.1 Epidemiology Studies 13

3.1.1 Mortality and Cancer Incidence Studies in Workers 13

3.1.2 Hormone Study in Male Workers 17

3.1.3 Occupational Study on Episodes of Care 18

3.1.4 3M Medical Surveillance Studies 20

3.1.4.1 Antwerp and Decatur Plants–Cross-Sectional 

and Longitudinal Studies 20

3.1.4.2 Cottage Grove Plant–Cross-sectional Studies of 

Clinical Chemistries and CCK 21

3.2 Metabolism and Pharmacokinetics 24

3.2.1 Metabolism and Pharmacokinetic Studies in Humans 24

3.2.2 Metabolism and Pharmacokinetic Studies in Non-Human Primates 24

3.2.3 Metabolism and Pharmacokinetic Studies in Adult Rats 26

3.2.3.1 Absorption Studies 26

3.2.3.1.1 Oral Exposure 26

3.2.3.1.2 Inhalation Exposure 27

3.2.3.1.3 Dermal Exposure 27

3.2.3.2 Serum Pharmacokinetic Parameters in Adult Rats 28

3.2.3.2.1 Oral and Intravenous Exposure in Sprague-Dawley Rats 28

3.2.3.2.2 Intravenous Exposure in Wistar Rats 30

3.2.3.3 Distribution Studies in Adult Rats 31

3.2.3.3.1 Oral Exposure 31

3.2.3.3.2 Intravenous Exposure 34

3.2.3.3.3 Intraperitoneal Exposure 34

3.2.3.4 Metabolism Studies in Adult Rats 36

3.2.3.5 Elimination Studies in Adult Rats 37

3.2.3.5.1 Enterohepatic Circulation 37

3.2.3.5.2 General Elimination Studies 37

3.2.3.5.2.1 Oral Exposure 37

3.2.3.5.2.2 Intravenous Exposure 37

3.2.3.5.3 Elimination Studies in the Pregnant Rat 38

3.2.3.5.4 Studies on the Mechanism of the Gender Difference 

in Elimination in Adult Rats 38

3.2.4 Metabolism and Pharmacokinetic Studies in Immature Rats 40

3.2.4.1 PFOA Levels During Pregnancy and Lactation 40

3.2.4.2 PFOA Levels in the Postweaning Rat 42

3.2.4.3 Serum and Tissue Distribution in Immature 

Wistar Rats Following Oral Exposure 43

3.2.5 Comparative Studies of Protein Binding in Humans, 

Non-Human Primates, and Rats 44

3.2.6 Metabolism and Pharmacokinetic Studies in Other Test Species 45

3.3 Acute Toxicity Studies in Animals 46

3.4 Mutagenicity Studies 47

3.5 Repeat Dose Studies in Animals 47

3.5.1 Subchronic Studies in Non-Human Primates 47
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3.5.2 Subchronic Studies in Rodents 52

3.5.3 Chronic Toxicity and Carcinogenicity Studies in Rats 55

3.6 Immunotoxicity Studies in Mice 59

3.7 Prenatal Developmental Toxicity Studies in Animals 60

3.8 Reproductive Toxicity Studies in Animals 64

3.8.1 F0 Generation 66

3.8.1.1 F0 Males 66

3.8.1.2 F0 Females 67

3.8.2 F1 Generation 68

3.8.2.1 F1 Males 68

3.8.2.2 F1 Females 69

3.8.3 F2 Generation 71

3.8.4 Conclusions 71

3.9 Mode of Action and Summary of Weight of Evidence 73

3.9.1 Epidemiology Studies 73

3.9.2 Metabolism and Pharmacokinetics 74

3.9.3 Mode of Action Analyses and Cancer Descriptor 75

3.9.3.1 Mode of Action Analysis of Liver Toxicity 

and Liver Adenomas in Rats 75

3.9.3.2 Human Relevance of the Rat PPAR"-agonist Induced 

Liver Toxicity and Liver Adenomas 80

3.9.3.2 Leydig Cell Adenomas in Rats 81

3.9.3.3 Pancreatic Acinar Cell Tumors in Rats 82

3.9.3.4 Cancer Descriptor 83

3.9.4 Toxicity in Adult Repeat-Dose Animal Studies 84

3.9.4.1 Non-Human Primates 84

3.9.4.2 Adult Male Rats 85

3.9.4.3 Adult Female Rats 87

3.9.4.4 Adult Mice 87

3.9.5 Developmental and Reproductive Toxicity in Animal Studies 88

4.0 Biomonitoring Data 90

4.1 Occupational Exposures 90

4.1.1 3M Occupational Data 90

4.1.2 DuPont Occupational Data 92

4.2 General Population Exposures 92

5.0 Risk Assessment 95

5.1 Selection of Endpoints 95

5.2 Use of Serum Levels as a Measure of Internal Dose in Humans 97

5.2.1 General Population 97

5.2.2 Workers 97

5.3 Calculation of MOEs Based on Non-Human Primate Studies 98

5.4 Calculation of MOEs Based on Adult Rat Studies 98

5.5 Calculation of MOEs Based on Rat Developmental Toxicity Studies 99

5.6 Uncertainties in the Risk Characterization 101

6.0 Overall Conclusions 103

7.0 References 105

Appendix A A-1
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Glossary of Abbreviations

ALT alanine aminotransferase, serum glutamyl pyruvic transaminase, SGPT

AP alkaline phosphatase

APFO ammonium salt of perfluorooctanoic acid

AST aspartate aminotransferase, serum glutamyl oxaloacetic transaminase,

SGOT

AUC area under the curve 

BMI body mass index

BUN blood urea nitrogen test 

CCG Clinical Care Groups software

CCK cholecystokinin-33 (human); cholecystokinin (rat)

CI confidence interval

Cl total body clearance

CLR renal clearance

ConA concanavalin A

DEHP di(2-ethylhexyl)phthalate

DHEAS dehydroepiandrosterone sulfate 

EGF epidermal growth factor 

FMPP Familial Male-limited Precocious Puberty

FID flame ionization detector

FSH follicle stimulating hormone

GD gestation day

GGT gamma glutamyl transferase

HDL high-density lipoprotein

17-HP 17 gamma-hydroxyprogesterone

HPLC/ESMS high performance liquid chromatography/electrospray tandem mass

spectrometry 

HRBC horse red blood cell

HSA human serum albumin

i.p. intraperitoneal

i.v. intravenous

Kd dissociation constant

LD lactation day

LDL low-density lipoprotein 

LH luteinizing hormone

LLOQ lower limit of quantitation

LOAEL lowest-observed-adverse-effect level 

LD50 lethal dose 50%

LPS lipopolysaccharide
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M570 N-methyl perfluorooctanesulfonamidoacetate

MOA mode of action

NMR nuclear magnetic resonance

NOAEL no-observed-adverse-effect level 

OAT organic anion transporter 

OPPT Office of Pollution Prevention and Toxics

PACT pancreatic acinar cell tumors 

PFHS perfluorohexanesulfonate

PFOA perfluorooctanoic acid

PFOS perfluorooctane sulfonate 

PFC plaque forming cell immune function test

PPAR" peroxisome proliferator-activated receptor "

ppb parts per billion

ppm parts per million

RREpC relative risk ratio for each episode of care

RSA rodent serum albumin

SEM standard error of the mean

SGOT serum glutamyl oxaloacetic transaminase, AST

SGPT serum glutamyl pyruvic transaminase, ALT

SHBG sex hormone-binding globulin

SIR standardized incidence ratio

SMR standardized mortality ratio 

t

1⁄2 or T 1⁄2 half life

TSH thyroid-stimulating hormone

Vd volume of distribution
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6

Executive Summary

As part of the effort by the Office of Pollution Prevention and Toxics (OPPT) to understand

health and environmental issues presented by fluorochemicals in the wake of unexpected

toxicological and bioaccumulation discoveries with respect to perfluorooctane sulfonates

(PFOS), OPPT has been investigating perfluorooctanoic acid (PFOA) and its salts. PFOA and

its salts are fully fluorinated organic compounds that can be produced synthetically or through

the degradation or metabolism of other fluorochemical products. PFOA is primarily used as a

reactive intermediate, while its salts are used as processing aids in the production of

fluoropolymers and fluoroelastomers and in other surfactant uses. PFOA and its salts are

persistent in the environment. Most of the toxicology studies have been conducted with the

ammonium salt of perfluorooctanoic acid, which is referred to as APFO in this report. 

Human Health Effects 

Epidemiological studies on the effects of PFOA in humans have been conducted on workers. 

Most of the studies were cross-sectional and focused primarily on males. Developmental

outcomes have not been studied. A retrospective cohort mortality study demonstrated a

statistically significant association between prostate cancer mortality and employment duration

in the chemical facility of a plant that manufactures PFOA. However, in an update to this study

in which more specific exposure measures were used, a significant association for prostate

cancer was not observed. Other mortality studies lacked adequate exposure data which could be

linked to health outcomes. A study which examined hormone levels in workers reported an

increase in estradiol levels in workers with the highest PFOA serum levels; however, these

results may have been confounded by body mass index. Cholesterol and triglyceride levels in

workers were positively associated with PFOA exposures, which is inconsistent with the

hypolipidemic effects observed in rat studies. A statistically significant positive association was

reported for PFOA and T3 levels in workers but not for any other thyroid hormones.

Little information is available concerning the pharmacokinetics of PFOA and its salts in humans. 

An ongoing 5-year, half-life study in 7 male and 2 female retired workers has suggested a mean

serum PFOA half-life of 4.37 years (range, 1.50 – 13.49 years). Gender differences in

elimination of PFOA have not been observed in humans based on the limited data available in

the half-life study in retired workers. Metabolism and pharmacokinetic studies in non-human

primates are limited to a study of 3 male and 3 female cynomolgus monkeys administered a

single i.v. dose of 10 mg/kg potassium PFOA. In male monkeys, the average serum half-life was

20.9 days. In female monkeys, the average serum half-life was 32.6 days. 

Studies in adult rats have shown that the ammonium salt of PFOA (APFO) is absorbed following

oral, inhalation and dermal exposure. Serum pharmacokinetic parameters and the distribution of

PFOA has been examined in the tissues of adult rats following administration by gavage and by

i.v. and i.p. injection. PFOA distributes primarily to the liver, serum, and kidney, and to a lesser

extent, other tissues of the body. It does not partition to the lipid fraction or adipose tissue. 

PFOA is not metabolized and there is evidence of enterohepatic circulation of the compound. 

The urine is the major route of excretion of PFOA in the female rat, while the urine and feces are

both main routes of excretion in male rats. 

There are gender differences in the elimination of PFOA in adult rats following administration

by gavage and by i.v. and i.p.injection. In female rats, following oral administration, estimates

of the serum half-life were dependent on dose and ranged from approximately 2.8 to 16 hours,
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while in male rats estimates of the serum half-life following oral administration were

independent of dose and ranged from approximately 138 to 202 hours. In female rats, 

elimination of PFOA appears to be biphasic with a fast phase and a slow phase. The rapid

excretion of PFOA by female rats is believed to be due to active renal tubular secretion (organic

acid transport system); this renal tubular secretion is believed to be hormonally controlled. 

Hormonal changes during pregnancy do not appear to cause a change in the rate of elimination in

rats. 

Several recent studies have been conducted to examine the kinetics of PFOA in the developing

Sprague-Dawley rat. These studies have shown that PFOA readily crosses the placenta and is

present in the breast milk of rats. During lactation and for the first several weeks after weaning,

the elimination of PFOA is similar in males and females pups. Between 4-5 weeks of age, the

elimination in male rats assumes the adult pattern and the gender difference becomes readily

apparent. Distribution studies in the postweaning rat have shown that PFOA is distributed

primarily to the serum, liver, and kidney.

In acute toxicity studies in animals, the oral LD50 values for CD rats were >500 mg/kg for males

and 250-500 mg/kg for females, and <1000 mg/kg for male and female Wistar rats. There was

no mortality following inhalation exposure of 18.6 mg/L for one hour in rats. The dermal LD50

in rabbits was determined to be greater than 2000 mg/kg. APFO is a primary ocular irritant in

rabbits, while the data regarding potential skin irritancy are conflicting.

APFO is not mutagenic. APFO did not induce mutation in either S. typhimurium or E. coli when

tested either with or without mammalian activation. APFO did not induce gene mutation when

tested with or without metabolic activation in the K-1 line of Chinese hamster ovary (CHO) cells

in culture. APFO did not induce chromosomal aberrations in human lymphocytes when tested

with and without metabolic activation up to cytotoxic concentrations. APFO was tested twice

for its ability to induce chromosomal aberrations in CHO cells. In the first assay, APFO induced

both chromosomal aberrations and polyploidy in both the presence and absence of metabolic

activation. In the second assay, no significant increases in chromosomal aberrations were

observed without activation. However, when tested with metabolic activation, APFO induced

significant increases in chromosomal aberrations and in polyploidy. APFO was negative in a

cell transformation assay in C3H 10T1⁄2 mouse embryo fibroblasts and in the mouse micronucleus

assay.

Repeat-dose studies have been conducted in non-human primates. In a 13-week study with

Rhesus monkeys, exposure to doses of 30 mg/kg-day or higher resulted in death. Clinical signs

of toxicity were noted at doses as low as 3 mg/kg-day. Unlike rodent studies, analyses of the

serum and liver levels did not reveal a gender difference in monkeys, but the sample size was

very small. In a 6-month study of male cynomolgus monkeys, there was a steep dose response

curve for mortality. Increases in liver weight were noted at doses as low as 3 mg/kg-day, but

there was no evidence of peroxisome proliferator-activated receptor alpha activity (PPAR").

Repeat-dose studies in rats and mice demonstrated that the liver is the primary target organ. Due

to gender differences in elimination, adult male rats exhibit effects at lower administered doses

than adult female rats. Dietary exposure to APFO for 90 days resulted in significant increases in

liver weight and hepatocellular hypertrophy in female rats at 1000 ppm (76.5 mg/kg-day) and in

male rats at doses as low as 100 ppm (5 mg/kg-day). Chronic dietary exposure of rats to 300

ppm(males, 14.2 mg/kg-day; females, 16.1 mg/kg-day) APFO for 2 years resulted in increased

liver weight, hepatocellular hypertrophy, hematological effects, and testicular masses in males;
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and reductions in body weight and hematological effects in females.

The carcinogenic potential of PFOA has been investigated in two dietary carcinogenicity studies

in rats. Under the conditions of these studies, there is some evidence that PFOA is carcinogenic,

inducing liver tumors, Leydig cell tumors (LCT), and pancreatic acinar cell tumors (PACT) in

male Sprague-Dawley rats. The evidence for mammary fibroadenomas in the female rats is

equivocal since the incidences were comparable to some historical background incidences. 

There is sufficient evidence to indicate that PFOA is a PPAR"-agonist and that the liver

carcinogenicity (and toxicity) of PFOA is mediated by binding to the PPAR" in the liver. A

mode of action analysis has demonstrated that the hepatic effects are due to PPAR"-agonism,

and that this mode of action is unlikely to occur in humans. The mode(s) of action for the

Leydig cell and pancreatic acinar cell tumors have been investigated, but there is insufficient

evidence to link these modes of action to PPAR". The LCT and PACT induced in the rat by

PFOA probably do not represent a significant cancer hazard for humans because of quantitative

differences in the expressions of LH and CCKA receptors and of other toxicodynamic differences

between the rat and the human. Based on no adequate human studies and uncertain relevance of

the tumors from the rat studies, PFOA may be best described as “suggestive evidence of

carcinogenicty, but not sufficient to assess human carcinogenic potential” under the draft 1999

EPA Guidelines for Carcinogen Risk Assessment.

PFOA appears to be immunotoxic in mice. Feeding C57Bl/6 mice a diet containing 0.02%

PFOA resulted in adverse effects to both the thymus and spleen. In addition, this feeding

regimen resulted in suppression of the specific humoral immune response to horse red blood

cells, and suppression of splenic lymphocyte proliferation. The suppressed mice recovered their

ability to generate a humoral immune response when they were fed a diet devoid of PFOA. 

Studies using transgenic mice showed that the PPAR" was involved in causing the adverse

effects to the immune system.

There was no evidence of prenatal developmental toxicity in rats after oral exposure to doses as

high as 150 mg/kg-day. Maternal toxicity was seen at 100 mg/kg-day. In a rabbit oral prenatal

developmental toxicity study there was a significant increase in skeletal variations after exposure

to 50 mg/kg-day APFO. There was no evidence of maternal toxicity at 50 mg/kg-day, the

highest dose tested. 

A variety of endpoints were evaluated throughout different lifestages in a two-generation

reproductive toxicity study in rats exposed to 0, 1, 3, 10, and 30 mg/kg/day APFO. In that study,

a reduction in F1 pup mean body weight on a litter basis was observed during lactation (sexes

combined) in the 30 mg/kg-day group. F1 male pups in the 10 and 30 mg/kg-day groups

exhibited a significant reduction in body weight gain during days 8-50 postweaning, and body

weights were significantly reduced in the 10 mg/kg-day group beginning on postweaning day 36,

and in the 30 mg/kg-day group beginning on postweaning day 8. F1 female pups in the 30

mg/kg-day group exhibited a significant reduction in body weight gain on days 1-15

postweaning, and in body weights beginning on day 8 postweaning. Reproductive indices were

not affected in the F1 animals. There was a significant increase in mortality mainly during the

first few days after weaning, and a significant delay in the timing of sexual maturation for F1

male and female pups in the 30 mg/kg-day group. No effects were observed in the F2 pups. 

However, it should be noted that the F2 pups were sacrificed at weaning, and thus it was not

possible to ascertain the potential post-weaning effects that were noted in the F1 generation. 

Adult systemic toxicity consisted of reductions in body weight in both the F0 and F1 animals. 
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Human Biomonitoring

While the environmental concentrations and pathways of human exposure to PFOA and its salts

are unknown, there are data on PFOA serum levels in workers and the general population. 

PFOA has been measured in the serum of workers occupationally exposed to perfluorinated

chemicals for many years. PFOA has also been detected recently in the serum of the general

U.S. population, but at much lower levels than those reported in occupational biomonitoring

studies. Individual blood serum samples from 3 separate non-occupational cohorts have been

analyzed for PFOA. Cohorts of adults (n= 645) and children (n=598) from various geographic

areas of the U.S. and an elderly cohort from Seattle (n = 238) have indicated consistent mean

serum levels of PFOA (approximately 5 ng/ml or 5 ppb). The reports indicate that serum levels

for most of the individuals in these samples are below 10 ng/ml; however, some of the levels are

as high as 56 ng/ml, indicating that a small number of individuals are being exposed at higher

concentrations than the rest of the general population.

Risk Assessment and Uncertainties

A margin of exposure (MOE) approach can be used to describe the potential for human health

effects associated with exposure to a chemical. The MOE is calculated as the ratio of the

NOAEL or LOAEL for a specific endpoint to the estimated human exposure level. The specific

endpoint may be from an epidemiology study or an animal toxicology study. The MOE does not

provide an estimate of population risk, but simply describes the relative “distance” between the

exposure level and the NOAEL or LOAEL. In this risk assessment there is no information on

the sources or pathways of human exposure. However, serum levels of PFOA, which are a

measure of cumulative exposure, were available from human biomonitoring studies. In addition,

serum levels of PFOA were available for many of the animal toxicology studies or there was

sufficient pharmacokinetic information to estimate serum levels. Thus, in this assessment

internal doses from animal and human studies were compared; this is somewhat analogous to a

MOE approach which uses external exposure estimates.

The results of existing epidemiology studies are not adequate for use in quantitative risk

assessment, and therefore the analysis was restricted to endpoints in the animal toxicology

studies. MOEs were calculated for the general U.S. population. Although some serum level

data were available for workers, the data were not adequate to calculate MOEs for occupational

exposures. In general, the mean serum levels following occupational exposures appear to be

orders of magnitude higher than observed in the general population. Thus, MOEs for workers

would be expected to be much less than for the general population. 

A variety of endpoints from the animal toxicology studies were used to calculate MOEs for this

draft risk assessment. The endpoints encompassed different species, gender and life stages. For

this draft assessment, specific recommendations on the most appropriate

endpoint/lifestage/species/gender have not been made; rather, all have been presented to provide

transparency.

For adults, two sets of MOEs were calculated based on the toxicology studies in non-human

primates and rats. First, MOEs were calculated from the cynomolgus monkey study and are 

based on increased liver weight and possible mortality. The MOE using the geometric mean for

the human serum level is 16,739 (8,191 for the 90th percentile). Second, MOEs were calculated

from the adult rat studies and are based on reductions in body weight. MOEs were calculated

separately for the female and male rat due to the gender differences in pharmacokinetics in this
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species. MOEs were calculated by dividing the AUC in the adult female rat by the AUC for the

adult humans which is 398 (195 for the 90th percentile) and by dividing the AUC for the adult

male rat by the AUC for the adult humans which is 9158 (4481 for the 90th percentile).

MOEs were calculated for the developmental effects in the two-generation reproductive toxicity

study in rats. These effects were observed at various times during the maturation of the F1 pups. 

For both F1 males and females there were reductions in body weight during lactation; significant

increases in mortality during the first few days after weaning; and significant delays in the

timing of sexual maturation. Mean body weights were also significantly reduced in the time

period prior to sexual maturation in both the F1 males and females. The critical period of

exposure for each of the effects is not known. For example, it is not known whether prenatal

and/or lactational exposure is important for the reduced body weight that was observed during

lactation. Similarly, it is not known whether the reduced body weight, mortality, or delayed

sexual maturation that occurred during the postweaning period are due to prenatal, lactational,

and/or postweaning exposures. Ideally, MOEs would be calculated for each of these exposure

periods; however MOEs were not calculated for the lactation period due to uncertainties in

pharmacokinetics. 

For the prenatal period, MOEs were calculated for the pregnant human female. MOEs were not

calculated for the fetus since there is no information on human serum levels in fetuses. MOEs

were calculated using both Cmax and AUC; the MOE based on Cmax is 3,095 (1548 for the 90th

percentile) and the MOE based on the AUC is 823 (412 for the 90th percentile). 

For the postweaning period, MOEs were calculated for several endpoints including reductions in

body weight, mortality and delayed sexual maturation. These MOEs were based on the

geometric mean for children and range from 10,484 - 78,546 (the range using the 90th is 6,044 -

45,279).

This assessment has provided a range of MOEs for several life stages. Several uncertainties

have been discussed in a qualitative fashion in the assessment, which highlight the need to

interpret the MOEs with caution. For example, MOEs were not calculated for the lactation

period due to insufficient data, although this may represent an important exposure period. 

Similarly, the biomonitoring data for the children are from samples collected in 1994 and may

not be representative of current children’s serum levels. Finally, there is some uncertainty

associated with the determination of the adequacy of a specific MOE in protecting human health

in the present context. Traditionally, MOEs are calculated from administered dose levels and

estimates of human exposure. In this assessment, the MOEs were calculated from internal dose

metrics in animals and humans. While use of internal dose metrics reduces many uncertainties

pertaining to exposure, there is little experience or guidance on the factors that should be

considered in making judgements about the level of concern associated with a given MOE. 

Approaches that are used for conventional MOEs, if applied unchanged, indicate that among the

populations of interest some individuals are highly exposed, for reasons not understood at this

time. However, if conventional approaches for determining levels of concern are not appropriate

for MOEs based on internal dose metrics, then this conclusion would have to be re-evaluated as

the understanding of this question evolves. 
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1.0 Scope of the Assessment

As part of the effort by the Office of Pollution Prevention and Toxics (OPPT) to understand

health and environmental issues presented by fluorochemicals in the wake of unexpected

toxicological and bioaccumulation discoveries with respect to perfluorooctane sulfonates

(PFOS), OPPT has been investigating perfluorooctanoic acid and its salts (PFOA). PFOA and

its salts are fully fluorinated organic compounds that can be produced synthetically or through

the degradation or metabolism of other fluorochemical products. PFOA is primarily used as a

reactive intermediate, while its salts are used as processing aids in the production of

fluoropolymers and fluoroelastomers and in other surfactant uses. 

OPPT released a preliminary Draft Hazard Assessment of Perfluorooctanoic Acid and Its Salts,

dated February 20, 2002, on March 28, 2002, and issued a minor correction to that document on

April 15, 2002. That draft assessment indicated that PFOA and its salts are persistent in the

environment and in humans with a half life of years. The assessment noted the potential

systemic toxicity and carcinogenicity associated with the ammonium salt of PFOA (APFO),

which has been the focus of the animal toxicology studies, and observed that blood monitoring

data suggested widespread exposure to the general population, albeit at low levels. The Agency

has since received considerable additional animal toxicology data on APFO that suggest a

potential for developmental/reproductive toxicity and immunotoxicity, and additional human

biomonitoring data that indicate low level exposures to the general population that cannot be

explained at this time. 

On September 27, 2002, the Director of OPPT issued a memorandum announcing that OPPT

would initiate a priority review to determine whether PFOA and its salts meets the criteria for

action under section 4(f) of the Toxic Substances Control Act. As part of the priority review, the

hazard assessment was revised and released on September 30, 2002. Another revision was then

released November 4, 2002. In addition, OPPT released a Preliminary Risk Assessmment of the

Developmental Toxicity of PFOA and Its Salts, dated March 4, 2003. Although there is a wide

range of toxicological endpoints associated with exposure to APFO, the initial scope of the

preliminary risk assessment focused only on the endpoints that are included in section 4(f); these

include cancer, mutations, and birth defects. OPPT did not include gene mutations in the

preliminary risk assessment since APFO is not known to be mutagenic. In addition, APFO is a

peroxisome proliferator-activated receptor " (PPAR") agonist and through this mode of action

could lead to the formation of liver tumors in rodents. The relevance of this mode of action for

humans is currently under scientific debate, and the Agency was engaged in activities to resolve

this issue. Therefore, the preliminary risk assessment was narrowly restricted to examine only

the potential risks of developmental toxicity.

Since the release of the preliminary risk assessment, several pieces of information have become

available. Several uncertainties regarding the pharmacokinetics of PFOA during development

were discussed in the preliminary risk assessment, and studies have now been conducted to

address these uncertainties. In addition, activities aimed at addressing the uncertainties

associated with the PPAR"-agonist mode of action have made significant progress, and a draft

Proposed OPPTS Science Policy paper was presented to the FIFRA Science Advisory Panel on

December 9, 2003. This new information has enabled OPPT to extend the assessment activities

such that this draft Risk Assessment of the Potential Human Health Effects Associated With

Exposure to PFOA and Its Salts considers all toxicological endpoints to the extent possible.

The relevant information pertaining to chemical properties, epidemiology, pharmacokinetics and
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12

metabolism, animal toxicology studies, and human exposure have been included in this draft risk

assessment. Other information pertaining to ecotoxicity, production and uses, fate and transport,

and environmental monitoring are avilable in the Draft Hazard Assessment of Perfluorooctanoic

Acid and Its Salts, dated November 4, 2002. 

2.0 Chemical Identity

Chemical Name: Perfluorooctanoic Acid

Molecular formula: C8 H F15 O2

Structural formula: F-CF2-CF2-CF2-CF2-CF2-CF2-CF2-C(=O)-X, 

The free acid and some common derivatives have the following CAS numbers:

The perfluorooctanoate anion does not have a specific CAS number. 

Free Acid (X = OM+; M = H) [335-67-1]

Ammonium Salt (X = OM+; M = NH4) [3825-26-1]

Sodium Salt (X = OM+; M = Na) [335-95-5]

Potassium Salt (X = OM+; M = K) [2395-00-8]

Silver Salt (X = OM+; M = Ag) [335-93-3]

Acid Fluoride (X = F) [335-66-0]

Methyl Ester (X = CH3) [376-27-2]

Ethyl Ester (X = CH2-CH3) [3108-24-5]

Synonyms: 1-Octanoic acid, 2,2,3,3,4,4,5,5,6,6,7,7,8,8,8-pentadecafluoro-PFOA

2.1 Physicochemical Properties

PFOA is a completely fluorinated organic acid. The typical structure has a linear chain of eight

carbon atoms. The physical chemical properties noted below are for the free acid, unless

otherwise stated. The data for the free acid, pentadecafluorooctanoic acid [335-67-1], is the

most complete. The reported vapor pressure of 10 mm Hg appears high for a low melting solid

when compared to other low melting solids (chloroacetic acid: solid; MP = 61 to 63 /C; BP =

189 /C; VP = 0.1 kPa [0.75 mm Hg] @ 20 /C; NIOSH, 1994), but is consistent with other

perfluorinated compounds with similar boiling points (perfluorobutanoic acid BP = 120 /C, VP

10 mm Hg @ 20 /C; Boit, 1975). Another explanation may be that the 10 mm vapor pressure

was measured at an elevated temperature (but the temperature inadvertently omitted), as

perfluorooctanoic acid is typically handled as a liquid at 65 /C (3M, 2001). The free acid is

expected to completely dissociate in water, leaving the anionic carboxylate in the water and the

perfluoroalkyl chain on the surface. In aqueous solutions, individual molecules of PFOA anion

loosely associate on the water surface and partition between the air / water interface. Several

reports note that PFOA salts self-associate at the surface, but with agitation they disperse and

micelles form at higher concentrations (Simister et al., 1992; Calfours and Stilbs, 1985; Edwards

et al., 1997). Water solubility has been reported for PFOA, but it is unclear whether these values

are for a microdispersion of micelles, rather than true solubility. Due to these same surfaceactive properties of PFOA, and the test protocol for the OECD shake flask method, PFOA is

anticipated to form multiple layers in octanol/water, much like those observed for PFOS. 

Therefore, an n-octanol/water partition coefficient cannot be determined.

The available physicochemical properties for the PFOA free acid are:
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Molecular weight: 414 (Boit, 1975)

Melting point: 45 – 50 /C (Boit, 1975)

Boiling point: 189 – 192 /C / 736 mm Hg (Boit, 1975)

Vapor pressure: 10 mm Hg @ 25 /C (approx.) (Exfluor, 1998)

Water solubility: 3.4 g/L (telomeric [MP = 34 /C ref. 0.01 - 0.02 mol/L ~4 - 8 g/L) (MSDS from

Merck, and Fischer, 2003)

pKa: 2.5 (Gilliland, 1992)

pH (1g/L): 2.6 (MSDS Merck)

The PFOA derivative of greatest concern and most wide spread use is the ammonium salt

(APFO; CAS No. 3825-26-1). The water solubility of APFO has been inconsistently reported. 

One 3M study reported the water solubility of APFO to be > 10%. It was noted in an earlier

study that at concentrations of 20 g/L, the solution “gelled” (Welter, 1979). These numbers

seem surprisingly low for a salt in light of Apollo Scientific selling a 31% aqueous solution of

APFO. Reported values for estimated partition coefficient (log Pow) of APFO do not agree. 

The anticipated formation of an emulsified layer between the octanol and water surface interface

would make determination of log Kow impossible. 

Determination of the vapor pressure of APFO is complicated. A vapor pressure of 7 x 10-5 mm

Hg at 20 /C has been reported for APFO; however, this appears to be too low for a material that

sublimates as the ammonium salt (Wolter, 1993). The ammonium salt begins to sublimate at

130 /C. As the temperature increases, 20% of the sample weight of APFO is lost by 169 /C. 

Other salts (Cs, K, Ag, Pb, Li) do not demonstrate similar weight loss until 237 /C or higher

(Lines and Sutcliffe, 1984). Decomposition of different salts produces perfluoroheptene (loss of

metal fluoride and carbon dioxide). This occurs at 320 /C for the sodium salt and at 250-290 /C

for the silver salt (Boit, 1975). 

3.0 Hazard Characterization

3.1 Epidemiology Studies

3M and Dupont have conducted several epidemiology and medical surveillance studies of the

workers at their plants in various cities of the U.S. No remarkable health effects that can be

directly attributed to PFOA exposure have been reported in fluorochemical production workers

described in the studies below. (Serum PFOA concentrations in workers volunteering in these

biomonitoring programs are presented in Table 24, Section 4.1.1 and Table 25, Section 4.1.2). 

3.1.1 Mortality and Cancer Incidence Studies in Workers

A retrospective cohort mortality study was performed on employees at the 3M Cottage Grove,

MN plant which produces APFO (Gilliland and Mandel, 1993). At this plant, APFO production

was limited to the Chemical Division. The cohort consisted of workers who had been employed

at the plant for at least 6 months between January 1947 and December 1983. Death certificates

of all of the workers were obtained to determine cause of death. There was almost complete

follow-up (99.5%) of all of the study participants. The exposure status of the workers was

categorized based on their job histories. If they had been employed for at least 1 month in the

Chemical Division, they were considered exposed. All others were considered to be not exposed

to PFOA. The number of months employed in the Chemical Division provided the cumulative

exposure measurements. Of the 3537 (2788 men and 749 women) employees who participated

in this study, 398 (348 men and 50 women) were deceased. Eleven of the 50 women and 148 of
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the 348 men worked in the Chemical Division, and therefore, were considered exposed to PFOA.

Standardized Mortality Ratios (SMRs), adjusted for age, sex, and race were calculated and

compared to U.S. and Minnesota white death rates for men. For women, only state rates were

available. The SMRs for males were stratified for 3 latency periods (10, 15, and 20 years) and 3

periods of duration of employment (5, 10, and 20 years). 

For all female employees, the SMRs for all causes and for all cancers were less than 1. The only

elevated (although not significant) SMR was for lymphopoietic cancer, and was based on only 3

deaths. When exposure status was considered, SMRs for all causes of death and for all cancers

were significantly lower than expected, based on the U.S. rates, for both the Chemical Division

workers and the other employees of the plant.

In all male workers at the plant, the SMRs were close to 1 for most of the causes of death when

compared to both the U.S. and the Minnesota death rates. When latency and duration of

employment were considered, there were no elevated SMRs. When employee deaths in the

Chemical Division were compared to Minnesota death rates, the SMR for prostate cancer for

workers in the Chemical Division was 2.03 (95% CI .55 - 4.59). This was based on 4 deaths

(1.97 expected). There was also a statistically significant (p = 0.03) association with length of

employment in the Chemical Division and prostate cancer mortality. Based on the results of

proportional hazard models, the relative risk for a 1-year increase in employment in the

Chemical Division was 1.13 (95% CI 1.01 to 1.27). It rose to 3.3 (95% CI 1.02 -10.6) for

workers employed in the Chemical Division for 10 years when compared to the other employees

in the plant. The SMR for workers not employed in the Chemical Division was less than

expected for prostate cancer (.58).

An update of this study was conducted to include the death experience of employees through

1997 (Alexander, 2001a). The cohort consisted of 3992 workers. The eligibility requirement

was increased to 1 year of employment at the Cottage Grove plant, and the exposure categories

were changed to be more specific. Workers were placed into 3 exposure groups based on job

history information: definite PFOA exposure (n = 492, jobs where cell generation, drying,

shipping and packaging of PFOA occurred throughout the history of the plant); probable PFOA

exposure (n = 1685, other chemical division jobs where exposure to PFOA was possible but with

lower or transient exposures); and not exposed to fluorochemicals (n = 1815, primarily nonchemical division jobs).

In this new cohort, 607 deaths were identified: 46 of these deaths were in the PFOA exposure

group, 267 in the probable exposure group, and 294 in the non-exposed group. When all

employees were compared to the state mortality rates, SMRs were less than 1 or only slightly

higher for all of the causes of death analyzed. None of the SMRs were statistically significant at

p = .05. The highest SMR reported was for bladder cancer (SMR = 1.31, 95% CI = 0.42 – 3.05).

Five deaths were observed (3.83 expected). 

A few SMRs were elevated for employees in the definite PFOA exposure group: 2 deaths from

cancer of the large intestine (SMR = 1.67, 95% CI = 0.02 – 6.02), 1 from pancreatic cancer

(SMR = 1.34, 95% CI = 0.03 – 7.42), and 1 from prostate cancer (SMR = 1.30, 95% CI = 0.03 –

7.20). In addition, employees in the definite PFOA exposure group were 2.5 times more likely

to experience cerebrovascular disease mortality (5 deaths observed, 1.94 expected; 95% CI =

0.84 – 6.03). 
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In the probable exposure group, 3 SMRs were elevated: cancer of the testis and other male

genital organs (SMR = 2.75, 95% CI = 0.07 – 15.3); pancreatic cancer (SMR = 1.24, 95% CI =

0.45 – 2.70); and malignant melanoma of the skin (SMR = 1.42, 95% CI = 0.17 – 5.11). Only 1,

6, and 2 cases were observed, respectively. The SMR for prostate cancer in this group was 0.86

(95% CI = 0.28 – 2.02) (n = 5). 

 

There were no notable excesses in SMRs in the non-exposed group, except for cancer of the

bladder and other urinary organs. Four cases were observed and only 1.89 were expected (95%

CI = 0.58 – 5.40).

It is difficult to interpret the results of the prostate cancer deaths between the first study and the

update because the exposure categories were modified in the update. Only 1 death was reported

in the definite exposure group and 5 were observed in the probable exposure group. All of these

deaths would have been placed in the chemical plant employees exposure group in the first

study. The number of years that these employees worked at the plant and/or were exposed to

PFOA was not reported. This is important because even 1 prostate cancer death in the definite

PFOA exposure group resulted in an elevated SMR for the group. Therefore, if any of the

employees’ exposures were misclassified, the results of the analysis could be altered

significantly. This issue has become more apparent, given the results of a biomonitoring study

that took place at the Cottage Grove plant in 2000 in which PFOA concentrations were not

correlated with years worked in the Chemical Division but instead were associated with the

specific area of the plant where APFO was produced (Olsen, et al., 2003f). 

The excess mortality in cerebrovascular disease noted in employees in the definite exposure

group was further analyzed based on number of years of employment at the plant. Three of the 5

deaths occurred in workers who were employed in jobs with definite PFOA exposure for more

than 5 years but less than 10 years (SMR = 15.03, 95% CI = 3.02 – 43.91). The other 2 occurred

in employees with less than 1 year of definite exposure. The SMR was 6.9 (95% CI = 1.39 –

20.24) for employees with greater than 5 years of definite PFOA exposure. In order to confirm

that the results regarding cerebrovascular disease were not an artifact of death certificate coding,

regional mortality rates were used for the reference population. The results did not change. 

When these deaths were further analyzed by cumulative exposure (time-weighted according to

exposure category), workers with 27 years of exposure in probable PFOA exposed jobs or those

with 9 years of definite PFOA exposure were 3.3 times more likely to die of cerebrovascular

disease than the general population. A dose-response relationship was not observed with years

of exposure. 

It is difficult to compare the results of the first and second mortality studies at the Cottage Grove

plant since the exposure categories were modified. Although the potential for exposure

misclassification was certainly more likely in the first study, it may still have occurred in the

update as well. It is difficult to judge the reliability of the exposure categories that were defined

without measured exposures. Although serum PFOA measurements were considered in the

exposure matrix developed for the update, they were not directly used. In the second study, the

chemical plant employees were sub-divided into PFOA-exposed groups, and the film plant

employees essentially remained in the “non-exposed” group. This was an effort to more

accurately classify exposures; however, these new categories do not take into account duration

of exposure or length of employment. Another limitation to this study is that 17 death

certificates were not located for deceased employees and therefore were not included in the

study. The inclusion or exclusion of these deaths could change the analyses for the causes of

death that had a small number of cases. Follow-up of worker mortality at Cottage Grove (and
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16

Decatur) needs to continue. Although there were more than 200 additional deaths included in

this analysis, it is a small number and the cohort is still relatively young. Given the results of

studies on fluorochemicals in both animals and humans, further analysis is warranted. 

Limited data are available on mortality and cancer incidence at Dupont’s Washington Works

Plant in Parkersburg, WV. These studies were periodically undertaken as part of a medical

surveillance program at the plant. The most recent report is summarized here. Cancer incidence

for active employees was reported for the 1959-2001 time period and mortality data were

reported for active and retired employees for 1957 through 2000 (Dupont, 2003). No other data,

such as employee exposure information, lifestyle factors, employee demographics, or other

chemicals used at the plant, are available in this report.

In the cancer incidence report, cancer cases were identified through a combination of company

health and life insurance claims and company cancer and mortality registries. Standardized

incidence ratios (SIRs) were only calculated for those cancers for which 5 or more cases were

observed, which included 14 types of cancer. Two of those cancer types were elevated and

statistically significant (p = 0.05): bladder [SIR = 1.9; 95% CI (1.15-3.07)] and kidney and

urinary organs [SIR = 2.3 (95% CI = 1.36-3.65)]. All of the reported cases were male. Some

other cancer types with elevated SIRs but which were not statistically significant at p = 0.05

included myeloid leukemia (2.02), cancer of the larynx (1.77), multiple myeloma and

immunoproliferative (1.72), malignant melanoma of skin (1.3), testicular cancer (1.46), and

brain cancer (1.2). 

In the mortality report, when all causes of death were reported, SMRs, adjusted for age and

gender, were statistically significant (p = 0.05) for rheumatic heart disease (SMR = 3.55; 95%

CI, 1.14-8.30) and atherosclerosis and aneurysm (SMR = 1.98; 95% CI, 1.17-3.14). 

Two separate analyses of leukemia incidence were conducted prior to the above studies at

Dupont’s Washington Works plant (Walrath and Burke, 1989; Karns and Fayerweather, 1991). 

The initial study reported a statistically significant (p = 0.10) elevated odds ratio (OR) of 2.1 for

leukemia incidence for male employees working at the plant from 1956-1989. Eight cases (all

male) of different types of leukemia were identified. The OR remained elevated when the

workers were divided into wage and salaried employees (2.2 and 2.0, respectively). In the

follow-up case-control study, four controls were selected from the plant for each case, matched

on gender, age and payroll status. Matched odds ratios were significantly elevated (p = 0.10) for

employees who had previously worked as custodians and engineers, 8.0 (90% CI, 1.1-60.0) and

7.9 (90% CI, 1.0 - 76.0), respectively and remained elevated (although not statistically

significant) for these same job categories within the plant (OR= 4.0 and 5.1, respectively). 

Matched OR were also reported based on the area of the plant where the cases worked; however,

no statistically significant (p = 0.10) OR were reported. 

The mortality data reported above do not show any statistically significant (p = 0.10) elevations

in leukemia deaths (all of the cases in the case-control study were dead at the time of the

mortality report), possibly because the number of cases was very small and divided among

different types of leukemias. The Washington Works data provide some insight as to where

more medical surveillance should be concentrated at this plant but provide little information

about the relationship of PFOA to mortality or cancer incidence since no exposure information,

use of other chemicals, or lifestyle information was collected on these employees.
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3.1.2 Hormone Study in Male Workers

Endocrine effects have been associated with PFOA exposure in animals; therefore, medical

surveillance data, including hormone testing, from male employees only of the Cottage Grove,

Minnesota plant were analyzed (Olsen et al., 1998a). PFOA serum levels were obtained for

volunteer workers in 1993 (n = 111) and 1995 (n = 80). Sixty-eight employees were common to

both sampling periods. In 1993, the range of PFOA was 0-80 ppm and 0-115 ppm in 1995 using

thermospray mass spectrophotometry assay. Eleven hormones were assayed from the serum

samples. They were: cortisol, dehydroepiandrosterone sulfate (DHEAS), estradiol, follicle

stimulating hormone (FSH), 17 gamma-hydroxyprogesterone (17-HP), free testosterone, total

testosterone, luteinizing hormone (LH), prolactin, thyroid-stimulating hormone (TSH) and sex

hormone-binding globulin (SHBG). Employees were placed into 4 exposure categories based on

their serum PFOA levels: 0-1 ppm, 1- < 10 ppm, 10- < 30 ppm, and >30 ppm. Statistical

methods used to compare PFOA levels and hormone values included: multivariable regression

analysis, ANOVA, and Pearson correlation coefficients. 

PFOA was not highly correlated with any of the hormones or with the following covariates: age,

alcohol consumption, body mass index (BMI), or cigarettes. Most of the employees had PFOA

serum levels less than 10 ppm. In 1993, only 12 employees had serum levels > 10 ppm, and 15

in 1995. However, these levels ranged from approximately 10 ppm to over 114 ppm. There

were only 4 employees in the >30 ppm PFOA group in 1993 and only 5 in 1995. Therefore, it is

likely that there was not enough power to detect differences in either of the highest categories. 

The mean age of the employees in the highest exposure category was the lowest in both 1993

and 1995 (33.3 years and 38.2 years, respectively). Although not significantly different from the

other categories, BMI was slightly higher in the highest PFOA category. 

When the mean values of the various hormones were compared by exposure categories, there

was a statistically significant (p = .01) elevation in prolactin in 1993 only for the 10 workers

whose serum levels were between 10 and 30 ppm compared to the lower 2 exposure categories. 

In addition, TSH was significantly (p = .002) elevated in the same exposure category for 1995

only (mean blood serum level was 2.9 ppm). However, mean TSH levels for the other exposure

categories, including the $ 30 ppm category, were all the same (1.7 ppm). In 1993, TSH was

elevated only in this same exposure category, as well; however did not reach statistical

significance (p = .09). 

Estradiol levels in the >30 ppm group in both years were 10% higher than the other PFOA

groups; however, the difference was not statistically significant (p <0.05). These results were

confounded by estradiol being highly correlated with BMI (r = .41, p < .001 in 1993, and r = .30,

p < .01 in 1995). In 1995, all 5 employees with PFOA levels > 30 ppm had BMIs > 28, although

this effect was not observed in 1993. The authors postulate that the study may not have been

sensitive enough to detect an association between PFOA and estradiol because measured serum

PFOA levels were likely below the observable effect levels suggested in animal studies (55 ppm

PFOA in the CD rat). Only 3 employees in this study had PFOA serum levels this high. They

also suggest that the higher estradiol levels in the highest exposure category could suggest a

threshold relationship between PFOA and estradiol. 

The authors did not report a negative association between PFOA serum levels and testosterone. 

There were no statistically significant trends (p < 0.05) noted for PFOA and either bound or free

testosterone. However, 17-HP, a precursor of testosterone, was highest in the >30 ppm PFOA

group in both 1993 and 1995. In 1995, PFOA was significantly associated with 17-HP in
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regression models adjusted for possible confounders. However, the authors state that this

association was based on the results of one employee (data were not provided in the report). 

Free testosterone was highly correlated with age in both 1993 and 1995 (r = -.48, p < .001; r = -

.40, p < .001, respectively). 

There are several design issues that should be noted when evaluating the results of this study. 

First, although there were 2 study years (1993 and 1995), the populations were not independent. 

Sixty-eight employees participated in both years. Second, there were 31 fewer employees who

participated in the study in 1995, thus reducing the power of the study. There were also very few

employees in either year with serum PFOA levels greater than 10 ppm. Third, the crosssectional design of the study does not allow for analysis of temporality of an association. Since

the half-life of PFOA is at least 1 year, the authors suggest that it is possible that there may be

some biological accommodation to the effects of PFOA. Fourth, only one sample was taken for

each hormone for each of the study years. In order to get more accurate measurements for some

of the hormones, pooled blood taken in a short time period should have been used for each

participant. Fifth, some of the associations that were measured in this study were done based on

the results of an earlier paper that linked PFOA with increased estradiol and decreased

testosterone levels. However, total serum organic fluorine was measured in that study instead of

PFOA, making it difficult to compare the results. Finally, there may have been some

measurement error of some of the confounding variables. 

3.1.3 Occupational Study on Episodes of Care

In order to gain additional insight into the effects of fluorochemical exposure on workers’ health,

an “episode of care” analysis was undertaken at the Decatur plant to screen for morbidity

outcomes that may be associated with long-term, high exposure to fluorochemicals (Olsen et al.,

2001g). An “episode of care” is a series of health care services provided from the start of a

particular disease or condition until solution or resolution of that problem. Episodes of care

were identified in employees’ health claims records using Clinical Care Groups (CCG) software. 

All inpatient and outpatient visits to health care providers, procedures, ancillary services and

prescription drugs used in the diagnosis, treatment, and management of over 400 diseases or

conditions were tracked. 

Episodes of care were analyzed for 652 chemical employees and 659 film plant employees who

worked at the Decatur plant for at least 1 year between January 1, 1993 and December 31, 1998. 

Based on work history records, employees were placed into different comparison groups: Group

A consisted of all film and chemical plant workers; Group B had employees who only worked in

either the film or chemical plant; Group C consisted of employees who worked in jobs with high

POSF exposures; and Group D had employees who worked in high exposures in the chemical

plant for 10 years or more prior to the onset of the study. Film plant employees were considered

to have little or no fluorochemical exposure, while chemical plant employees were assumed to

have the highest exposures. 

Ratios of observed to expected episodes of care were calculated for each plant. Expected

numbers were based on 3M’s employee population experience using indirect standardization

techniques. A ratio of the chemical plant’s observed to expected experience divided by the film

plant’s observed to expected experience was calculated to provide a relative risk ratio for each

episode of care (RREpC). For each RREpC, 95% confidence intervals were calculated. 

Episodes of care that were of greatest interest were those which had been reported in animal or

epidemiologic literature on PFOS and PFOA: liver and bladder cancer, endocrine disorders
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involving the thyroid gland and lipid metabolism, disorders of the liver and biliary tract, and

reproductive disorders. 

The only increased risk of episodes for these conditions of a priori interest were for neoplasms of

the male reproductive system and for the overall category of cancers and benign growths (which

included cancer of the male reproductive system). There was an increased risk of episodes for

the overall cancer category for all 4 comparison groups. The risk ratio was greatest in the group

of employees with the highest and longest exposures to fluorochemicals (RREpC = 1.6, 95% CI

= 1.2 – 2.1). Increased risk of episodes in long-time, high-exposure employees also was reported

for male reproductive cancers (RREpC = 9.7, 95% CI = 1.1 - 458). It should be noted that the

confidence interval is very wide for male reproductive cancers and the sub-category of prostate

cancer. Five episodes of care were observed for reproductive cancers in chemical plant

employees (1.8 expected), of which 4 were prostate cancers (RREpC = 8.2, 95% CI = 0.8 - 399). 

One episode of prostate cancer was observed in film plant employees (3.4 expected). This

finding should be noted because an excess in prostate cancer mortality was observed in the

Cottage Grove plant mortality study when there were only 2 exposure categories (chemical

division employees and non-chemical division employees). The update of the study sub-divided

the chemical plant employees and did not corroborate this finding when exposures were divided

into definitely exposed and probably exposed employees. 

There was an increased risk of episodes for neoplasms of the gastrointestinal tract in the high

exposure group (RREpC = 1.8, 95% CI = 1.2- 3.0) and the long-term employment, high

exposure group (RREpC = 2.9, 95% CI = 1.7 – 5.2). Most of the episodes were attributable to

benign colonic polyps. Similar numbers of episodes were reported in film and chemical plant

employees.

In the entire cohort, only 1 episode of care was reported for liver cancer (0.6 expected) and 1 for

bladder cancer (1.5 expected). Both occurred in film plant employees. Only 2 cases of cirrhosis

of the liver were observed (0.9 expected), both in the chemical plant. There was a greater risk of

lower urinary tract infections in chemical plant employees, but they were mostly due to recurring

episodes of care by the same employees. It is difficult to draw any conclusions about these

observations, given the small number of episodes reported. 

Chemical plant employees in the high exposure, long-term employment group were 2 1⁄2 times

more likely to seek care for disorders of the biliary tract than their counterparts in the film plant

(RREpC = 2.6, 95% CI = 1.2 - 5.5). Eighteen episodes of care were observed in chemical plant

employees and 14 in film plant workers. The sub-categories that influenced this observation

were episodes of cholelithiasis with acute cholecystitis and cholelithiasis with chronic or

unspecified cholecystitis. Most of the observed cases occurred in chemical plant employees.

Risk ratios of episodes of care for endocrine disorders, which included sub-categories of thyroid

disease, diabetes, hyperlipidemia, and other endocrine or nutritional disorders, were not elevated

in the comparison groups. Conditions which were not identified a priori but which excluded the

null hypothesis in the 95% confidence interval for the high exposure, long-term employment

group included: disorders of the pancreas, cystitis, and lower urinary tract infections.

The results of this study should only be used for hypothesis generation. Although the episode of

care design allowed for a direct comparison of workers with similar demographics but different

exposures, there are many limitations to this design. The limitations include: 1) episodes of care

are reported, not disease incidence, 2) the data are difficult to interpret because a large RREpC
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may not necessarily indicate high risk of incidence of disease, 3) many of the risk ratios for

episodes of care had very wide confidence intervals, thereby providing unstable results, 4) the

analysis was limited to 6 years, 5) the utilization of health care services may reflect local

medical practice patterns, 6) individuals may be counted more than once in the database because

they can be categorized under larger or smaller disease classifications, 7) episodes of care may

include the same individual several times, 8) not all employees were included in the database,

such as those on long-term disability, 9) the analysis may be limited by the software used, which

may misclassify episodes of care, 10) the software may assign 2 different diagnoses to the same

episode, and 11) certain services, such as lab procedures may not have been reported in the

database.

3.1.4 3M Medical Surveillance Studies

3.1.4.1 Antwerp and Decatur Plants–Cross-Sectional and Longitudinal Studies 

A cross-sectional analysis of the data from the 2000 medical surveillance program at the Decatur

and Antwerp plants was undertaken to determine if there were any associations between PFOA

and hematology, clinical chemistries, and hormonal parameters of volunteer employees (Olsen et

al., 2001e). The data were analyzed for all employees from both plant locations. Mean PFOA

serum levels were 1.03 ppm for all male employees at the Antwerp plant and 1.90 ppm for all

male employees at the Decatur plant. Male production employees at the Decatur plant had

significantly higher (p < .05) mean serum levels (2.34 ppm) than those at the Antwerp plant

(1.28 ppm). Non-production employees at both plants had mean levels below 1 ppm. PFOA

serum levels were higher than the PFOS serum values at both plants, especially the Decatur plant

where serum levels are higher overall. In addition, values for total organic fluorine were even

higher than the PFOA levels.

Multivariable regression analyses were conducted to adjust for possible confounders that may

affect the results of the clinical chemistry tests. The following variables were included: 

production job (yes or no), plant, age, BMI, cigarettes/day, drinks/day and years worked at the

plant. A positive significant association was reported between PFOA and cholesterol (p = .05)

and PFOA and triglycerides (p = .002). Age was also significant in both analyses. Alcohol

consumed per day was significant in the cholesterol model, while BMI and cigarettes smoked

per day was significant for triglycerides. When both PFOA and PFOS were included in the

analyses, neither reached statistical significance in the cholesterol model, while PFOA remained

significant (p = .02) in the triglycerides model. High-density lipoprotein (HDL) was negatively

associated with PFOA (p = .04) and remained significant (p = .04) when both PFOA and PFOS

were included in the model. A positive association (p = .01) between T3 and PFOA was also

observed and remained statistically significant (p = .05) when PFOS was included in the model. 

BMI, cigarettes/day, alcohol/day were also significant in the model. None of the other clinical

chemistry, thyroid or hematology measures were significantly associated with PFOA in the

regression model.

A longitudinal analysis of the above data and previous medical surveillance results was

performed to determine whether occupational exposure to fluorochemicals over time is related to

changes in clinical chemistry and lipid results in employees of the Antwerp and Decatur

facilities (Olsen et al., 2001f). The clinical chemistries included: cholesterol, HDL,

triglycerides, alkaline phosphatase, gamma glutamyl transferase (GGT), aspartate

aminotransferase (AST), alanine aminotransferase (ALT), total and direct bilirubin. Medical

surveillance data from 1995, 1997, and 2000 were analyzed using multivariable regression
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analysis. The plants were analyzed using 3 subcohorts that included those who participated in 2

or more medical exams between 1995 and 2000. A total of 175 male employees voluntarily

participated in the 2000 surveillance and at least one other. Only 41 employees were

participants in all 3 surveillance periods.

When mean serum PFOA levels were compared by surveillance year, PFOA levels in the

employees participating in medical surveillance at the Antwerp plant increased between 1994/95

and 1997 and then decreased slightly between 1997 and 2000. At the Decatur plant, PFOA

serum levels decreased between 1994/95 and 1997 and then increased between 1997 and 2000. 

When analyzed using mixed model multivariable regression and combining Antwerp and

Decatur employees, there was a statistically significant positive association between PFOA and

serum cholesterol (p = .0008) and triglycerides (p = .0002) over time. When analyzed by plant

and also by subcohort, these associations were limited to the Antwerp employees (p = .005) and,

in particular, the 21 Antwerp employees who participated in all 3 surveillance years (p = .001). 

However, the association between PFOA and triglycerides was also statistically significant 

(p = .02) for the subgroup in which employees participated in biomonitoring in 1994/95 and

2000. There was not a significant association between PFOA and triglycerides among Decatur

workers. There were no significant associations between PFOA and changes over time in HDL,

alkaline phosphatase, GGT, AST, ALT, total bilirubin, and direct bilirubin.

There are several limitations to the 2000 cross-sectional and longitudinal studies including: 

1) serum PFOA levels were significantly higher at the Decatur plant than at the Antwerp plant,

2) all participants were volunteers, 3) there were several consistent differences in clinical

chemistry profiles and demographics between employees of the Decatur and Antwerp plants

(Antwerp employees as compared to Decatur employees had lower PFOA serum levels, were

younger, had lower BMIs, worked fewer years, had higher alcohol consumption, higher mean

HDL and bilirubin values, lower mean triglyceride, alkaline phosphatase, GGT, AST, and ALT

values, and mean thyroid hormone values tended to be higher), 4) PFOS and other perfluorinated

chemicals are also present in these plants, 5) in the cross-sectional study, plant populations

cannot be compared because they were placed into quartiles based on PFOS serum distributions

only which were different for each subgroup and not applicable to PFOA, 6) only one

measurement at a certain point in time was collected for each clinical chemistry test, and 7)

PFOA serum levels overall have been increasing over time in these employees. In addition, in

the longitudinal study only a small number of employees participated in all 3 sampling periods

(24%), different labs and analytical techniques for PFOA were used each year, and female

employees could not be analyzed because of the small number of participants.

3.1.4.2 Cottage Grove Plant–Cross-sectional Studies of Clinical Chemistries and CCK

A voluntary medical surveillance program was offered to employees of the Cottage Grove,

Minnesota plant in 1993, 1995, and 1997 (n = 111, 80 and 74 employees, respectively) (Olsen et

al., 1998b, 2000). The clinical chemistry parameters (cholesterol, hepatic enzymes, and

lipoprotein levels) used in the longitudinal and cross-sectional studies of the Antwerp and

Decatur plants were also used in this study. In addition, in 1997 only, cholecystokinin-33 (CCK)

was also measured at the Cottage Grove plant. CCK levels were observed because certain

research has suggested that pancreas acinar cell adenomas seen in rats exposed to PFOA may be

the result of increased CCK levels (Obourn et al., 1997). 

Only male employees involved in PFOA production were included in the study. Sixty-eight

employees were common to the 1993 and 1995 sampling periods, 21 were common between
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22

1995 and 1997, and 17 participated in all three surveillance years. Cottage Grove has the highest

serum PFOA levels of the 3 plants studied.

Employees’ serum PFOA levels were stratified into 3 categories (<1, 1- <10, and $10 ppm),

chosen to provide a greater number of employees in the $ 10 ppm category. As employees’

mean serum PFOA levels increased, no statistically significant (p < 0.05) abnormal liver

function tests, hypolipidemia, or cholestasis were observed in any of the sampling years. 

Multivariable regression analyses controlling for potential confounders (age, alcohol

consumption, BMI, and cigarettes smoked) yielded similar results. The authors also reported

that renal function, blood glucose, and hematology measures were not associated with serum

PFOA levels; however, these data were not provided in the paper.

The mean CCK value reported for the 1997 sample was 28.5 pg/ml (range 8.8 - 86.7 pg/ml). 

The means in the 2 serum categories < 10 ppm were at least 50% higher than in the $ 10 ppm

category. A statistically significant (p = .03) negative association between mean CCK levels and

the 3 PFOA serum categories was observed. A scatter plot of the natural log of CCK and PFOA

shows that all but 2 CCK values are within the assay’s reference range of 0 - 80 pg/ml. Both of

these employees (CCK values of 80.5 and 86.7 pg/ml) had serum PFOA levels less than 10 ppm

(0.6 and 5.6 ppm, respectively). A multiple regression model of the natural log of CCK and

serum PFOA levels continued to display a negative association after adjusting for potential

confounders. 

The cross-sectional design is a limitation of this study. Only 17 subjects were common to all 3

sampling years. In addition, the medical surveillance program is a voluntary one. The

participation rate of eligible production employees decreased from approximately 70% in 1993

to 50% in 1997. Also, the laboratory reference range changed substantially for ALT in 1997. 

Finally, different analytical methods were used to measure serum PFOA. Serum PFOA was

determined by electrospray high-performance liquid chromatography/mass spectrometry in

1997, but by thermospray in 1993 and 1995. 

3M used data collected in their 2000 medical surveillance program to determine whether serum

PFOA levels greater than 5 ppm were associated with changes in hepatic, lipid and thyroid

function in workers (Olsen, et al., 2003f). Clinical chemistries, including thyroid function tests,

were performed for 131 male and 17 female workers at the plant. Serum samples were extracted

and quantitatively analyzed for PFOA using HPLC/EMSS. All of the samples were above the

lower limit of quantitation (LLOQ) (see Table 24 in Section 4.1.1 for biomonitoring data).

Fifteen percent (n = 20) of male employees had serum concentrations that exceeded 5 ppm, and 

none of the female employees were above 5 ppm. Number of years worked in the Chemical

Division of the plant were not correlated with PFOA serum measurements, but were correlated

with specific production areas. When the male employees were separated into 3 groups based on

serum PFOA levels (< 1 ppm, 1- 4.9 ppm, and >5 ppm), there were no statistically significant (p

< .05) differences in mean lipid and hepatic test results or in thyroid hormone levels, between the

3 groups both before and after adjusting for potential confounders (eg. BMI, smoking status, and

alcohol consumption). The elimination of employees receiving cholesterol-reducing drugs (n =

9) from the analysis did not alter these findings. Similar to findings at the Decatur and Antwerp

plants, triglyceride levels were higher in employees with the highest PFOA serum

concentrations, although not statistically significant (p < .05). In simple linear regression

analyses, a weak negative association between T4 and serum PFOA concentrations was reported

(p = .07). However, all of the serum PFOA concentrations were within the T4 reference range,
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23

the statistical association explained minimal variation in the model (r2

 = .03), there was no

increase in serum TSH or T3 levels, and no negative association between free T4 levels and

PFOA. No statistically significant (p < 0.05) associations between PFOA and clinical

chemistries or thyroid test results for the small group of female employees was reported. 

An earlier medical surveillance study on workers who were employed in the 1980’s was

conducted at the Cottage Grove plant; however, total serum fluorine was measured instead of

PFOA (Gilliland and Mandel, 1996). Based on animal studies that reported that animals exposed

to PFOA develop hepatomegaly and alterations in lipid metabolism, a cross-sectional,

occupational study was performed to determine if similar effects are present in workers exposed

to PFOA. In a PFOA production facility, 115 workers were studied to determine whether serum

PFOA affected their cholesterol, lipoproteins, and hepatic enzymes. Forty-eight workers who

were exposed to PFOA from 1985-1989 were included in the study (96% participation rate). 

Sixty-five employees who either volunteered or were asked to participate, were included in the

unexposed group. These employees were assumed to have little or no PFOA exposure based on

their job description. However, when serum levels were analyzed, it was noted that this group of

workers had PFOA levels much greater than the general population. Therefore, instead of job

categories, total serum fluorine was used to classify workers into exposure groups.

Total serum fluorine was used as a surrogate measure for PFOA. Serum PFOA was not

measured, due to the cost of analyzing the samples. Blood samples were analyzed for total

serum fluorine, serum glutamyl oxaloacetic transaminase (SGOT or AST), serum glutamyl

pyruvic transaminase (SGPT or ALT), GGT, cholesterol, low-density lipoproteins (LDL), and

HDL. All of the participants were placed into five categories of total serum fluorine levels: <1

ppm, 1-3 ppm, >3 - 10 ppm, >10 - 15 ppm, and > 15 ppm. The range of the serum fluorine

values was 0 to 26 ppm (mean 3.3 ppm). Approximately half of the workers fell into the > 1 - 3

ppm category, while 23 had serum levels < 1 ppm and 11 had levels > 10 ppm.

There were no significant differences between exposure categories when analyzed using

univariate analyses for cholesterol, LDL, and HDL. In the multivariate analysis, there was not a

significant association between total serum fluorine and cholesterol or LDL after adjusting for

alcohol consumption, age, BMI, and cigarette smoking. There were no statistically significant

differences among the exposure categories of total serum fluorine for AST, ALT and GGT. 

However, increases in AST and ALT occurred with increasing total serum fluorine levels in

obese workers (BMI = 35 kg/m2

). This result was not observed when PFOA was measured

directly in serum of workers in 1993, 1995, or 1997 surveillance data of employees of the

Cottage Grove plant (Olsen et al., 2000). 

Since PFOA was not measured directly and there is no exposure information provided on the

employees (e.g. length of employment/exposure), the results of the study provide limited

information. The authors state that no adverse clinical outcomes related to PFOA exposure have

been observed in these employees; however, it is not clear that there has been follow-up of

former employees. In addition, the range of results reported for the liver enzymes were fairly

wide for many of the exposure categories, indicating variability in the results. Given that only

one sample was taken from each employee, this is not surprising. It would be much more

helpful to have several samples taken over time to ensure their reliability. It also would have

been interesting to compare the results of the workers who were known to be exposed to PFOA

to those who were originally thought not to be exposed to see if there were any differences

among the employees in these groups. There were more of the “unexposed” employees (n = 65)

participating in the study than those who worked in PFOA production (n = 48).
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24

3.2 Metabolism and Pharmacokinetics

3.2.1 Metabolism and Pharmacokinetic Studies in Humans

Little is known about the metabolism and pharmacokinetics of PFOA in humans. One report

notes the presence of PFOA in the cord blood of some pregnant workers suggesting that PFOA

can cross the placenta (U.S. EPA, 2003). In addition, there are limited data on the half-life of

PFOA in humans. With the exception of a 1980 study in which total organic fluorine in blood

serum was measured in one worker, no other data were available until June 2000 (Ubel et al.,

1980). A half-life study on 27 retirees from the Decatur and Cottage Grove 3M plants was

undertaken, in which serum samples were drawn every 6 months over a 5-year period. Two

interim reports describing the results thus far have been submitted (Burris et al., 2000, 2002). 

The first interim report suggested a median serum half-life of PFOA of 344 days, with a range of

109 to 1308 days. The two highest half-life calculations were for the 2 female retirees who

participated in this study (654 and 1308 days).

There were several limitations to this first analysis including: 1) the limited data available and

the range of serum PFOA levels measured; 2) serum was analyzed after each collection period

with only one measurement per time period on different days using slightly different analytical

techniques; and 3) the reference material purity was not determined until after the first 3 samples

had been analyzed. 

An effort was made to minimize experimental error, including systematic and random error in

the analytical method. Serum samples were collected from 9 of the original 27 subjects over 4

time periods spanning 180 days, measured in triplicate with all time points from each subject

analyzed in the same analytical run. This would allow for statistical evaluation of the precision

of the measurement and assure that all systematic error inherent in the assay equally affected

each sample used for half-life determination. Single serum measurements were made on

samples of the remaining 18 retirees, but were not included by the investigators in the analysis

because triplicate analyses of all time points were not conducted.

Of the 9 retirees included in this analysis, there were 7 males and 2 females, all from the Decatur

plant. The average age of the retirees was 61 years, the mean number of years worked at

Decatur was 27.7 years, and the average number of months retired from the plant at study

initiation was 18.9. The average BMI of this group was 27.9 (range 22.5-33, SD = 3.6). The

mean PFOA value at study initiation was 0.72 ppm (range 0.06 – 1.84 ppm, SD = 0.64).

The mean serum half-life for PFOA was 4.37 years (range 1.50 – 13.49 years, SD = 3.53). Only

1 employee had a half-life value that exceeded 4.3 years. The 2 females had values of 3.1 and

3.9 years. Age, BMI, number of years worked or years since retirement were not significant

predictors of serum half-lives in multivariable regression analyses.

This analysis has attempted to reduce experimental error in the determination of a half-life for

PFOA. However, two issues should be noted. First, the effect of continual non-occupational,

low-level exposure on the half-life is unknown. Second, systematic error of the analytical

method could be as high as +/- 20% and still satisfy the data quality criteria. 

3.2.2 Metabolism and Pharmacokinetic Studies in Non-Human Primates

Noker and Gorman (2003) administered a single intravenous (i.v.) dose of 10 mg/kg potassium
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25

PFOA to 3 male and 3 female cynomolgus monkeys. The animals were approximately 3-4 years

of age at the start of the study. Each monkey was examined shortly after dosing for clinical

signs of toxicity; all animals were observed twice daily for signs of mortality/moribundity. 

Additional clinical observations were performed on the days blood was collected. Each animal

was weighed on days 1, 4, 7, 14, 21 and 28. Urine and feces were collected on days 0, 1, 2, 7,

14, 21, and 28. Blood was collected at 0, 0.5, 2, 4, 8 and 24 hours and on days 2, 4, 7, 11, 14, 21,

28, 57, 79, 87, and 123. Serum and urine samples were analyzed by HPLC/MS/MS. 

No deaths that could be attributed to administration of the test article occurred during the study.

One male was euthanized on day 79 because he had developed repeated episodes of selfmutilation. These episodes do not appear to be related to PFOA administration. No adverse

clinical signs were noted in any of the other monkeys during the course of the study. Body

weights of the treated animals did not change between days 1 and 28.

At 0.5 hours after dosing, serum concentration in males and females were similar, and ranged

from 91,130 to 96,660 ng/ml (ppb) in males and from 88,940 to 96,400 ng/ml in females. Serum

concentrations declined slowly, and the levels decreased faster in two of the male monkeys than

in the third male monkey and all three female monkeys. By day 123, PFOA concentrations were

at or slightly above the limit of quantitation (20 ng/ml) in the two surviving males and between

885 and 4701 ng/ml in the three females.

The pharmacokinetic parameters that were calculated from the serum concentrations of PFOA

are presented in Table 1. The estimated values for half-life and total body clearance indicated

that two of the three male monkeys eliminated PFOA at a faster rate than did the female

monkeys. The volume of distribution of PFOA at steady state was similar for both sexes. Male

#2052 appeared to behave more like the females in his pharmacokinetic parameters than either

of the other two males in the study. One explanation is the stress of the experiment caused this

monkey to release high levels of cortisol that could have affected carbohydrate, protein, and/or

lipid metabolism, caused a shift in electrolyte and water balance, or increased plasma proteins.

This animal was also observed to have periods of severe self-mutilation and euthanized for this

reason on day 79.
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26

Table 1

Pharmacokinetic Parameters Calculated from Serum Concentrations of PFOA in 

Cynomolgus Monkeys

Parameter

Males Females

2052 2054 2211 Mean S.D. 2058 2059 2061 Mean S.D.

Cmax

(:g/mL)a 101 98.4 91.6 97.0 4.9 105 93.0 103 100 6.0

t

1⁄2 (day)

b 35.3 13.7 13.6 20.9 12.5 26.8 41.7 29.3 32.6 8.0

AUC0-last

(:g•day/mL)c 1999 569 633 1067 808 3083 2314 1056 2151 1023

AUC0-infinity

(:g•day/mL)d 2501 571 634 1235 1097 3224 2560 1094 2293 1090

Cl (mL/day/kg)e 4.0 17.5 15.8 12.4 7.7 3.1 3.9 9.1 5.4 3.3

Vdss (mL/kg)f 192 168 184 181 12 133 190 270 198 69

a Maximum concentration in serum

b Half-life of the terminal elimination phase

c Area under the serum concentration versus time curve calculated from 0 to the last time point

d Area under the serum concentration versus time curve calculated from 0 to infinity

e Total body clearance

f Volume of distribution at steady state

There were no obvious sex differences in the urinary excretion of PFOA. The rate of urinary

excretion by the monkeys was slow. Less than 20% of the administered dose was excreted in the

urine of the male and female monkeys within the first 48 hours after dosing, and as much as 0.1 -

1% of the administered dose was excreted in the urine on day 28. 

3.2.3 Metabolism and Pharmacokinetic Studies in Adult Rats

3.2.3.1 Absorption Studies 

Studies in rats have shown that PFOA is absorbed following oral, inhalation, and dermal

exposures.

3.2.3.1.1 Oral Exposure

Gibson and Johnson (1979) administered a single dose of 11.0 mg/kg 14C-PFOA by gavage to

groups of 3 male 10-week old CD rats. Twenty-four hours after administration, at least 93% of

the total carbon-14 was absorbed; the elimination half-life of carbon-14 from the plasma was 4.8

days.
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27

Ophaug and Singer (1980) administered 2 ml of an aqueous solution of 2 mg PFOA to female

Holtzman rats. Ionic fluoride, nonionic fluorine, and total fluorine were measured. Within 4.5

hr, 37% of the total fluorine in the administered dose was recovered in the urine. The quantity of

nonionic fluorine recovered in the urine increased to 61%, 76% and 89% at 8, 24 and 96 hr,

respectively, after administration. Within 4.5 hr, serum from treated rats had a nonionic fluorine

level of 13.6 ppm. The nonionic fluorine level in the serum decreased to 11.2 ppm at 8 hr, 0.35

ppm at 24 hr, and 0.08 ppm at 96 hr. Despite the large decrease in nonionic fluorine in the

serum, the ionic fluoride level was only 0.03 ppm and remained at that level throughout the

experiment. Prior to administration of PFOA, the ionic and nonionic fluorine background levels

in serum were 0.032 and 0.07 ppm, respectively. The authors concluded that PFOA is rapidly

absorbed from the gastrointestinal tract and then rapidly cleared from the serum.

3.2.3.1.2 Inhalation Exposure

Hinderliter (2003) measured the serum concentrations of PFOA following single and repeated

inhalation exposures in Sprague-Dawley rats. For the single exposure study, male and female

rats (3/sex/group) were exposed to a single nose-only exposure of an aerosol of 0, 1, 10, or 25

mg/m3

 PFOA. Preliminary range-finding studies demonstrated that aerosol sizes were 1.8 - 2.0

:m mass median aerodynamic diameter (MMAD) with geometric standard deviations ranging

from 1.9 - 2.1 :m. Blood samples were collected pre-exposure, at 0.5, 1, 3 and 6 hours during

exposure, and at 1, 3, 6, 12, 18 and 24 hours after exposure. Plasma was analyzed by LC-MS. 

PFOA plasma concentrations were proportional to the inhalation exposure concentrations. The

male Cmax values were approximately 2-3 times higher than the female Cmax. The female Cmax

occurred approximately one hour after the exposure period, while the male Cmax occurred from

the end of the exposure period up to six hours after exposure. In females, the elimination of

PFOA was rapid at all exposure levels, and by 12 hours after exposure the plasma levels had

dropped below the analytical limit of quantitation (0.1 :g/ml). In males, the plasma elimination

was much slower, and at 24 hours after exposure, the plasma concentrations were approximately

90% of the peak concentrations at all exposure levels.

In the repeated dose study, Hinderliter (2003) exposed male and female rats (5/sex/group) to the

same aerosol concentrations of PFOA for 6 hrs/day, 5 days/week for 3 weeks. Blood was

collected immediately before and after the daily exposure period three days per week. The

aerosol sizes were 1.3 - 1.9 :m MMAD with geometric standard deviations of 1.5 - 2.1. PFOA

plasma concentrations were proportional to the inhalation exposure concentrations, and repeated

exposures produced little plasma carryover in females, but significant carryover in males. Male

rats reached steady state plasma levels by three weeks with plasma concentrations of 8, 21, and

36 :g/ml for the 1, 10 and 25 mg/m3

 groups, respectively. In females, the post-exposure plasma

levels were 1, 2, and 4 :g/ml for the 1, 10, and 25 mg/m3 groups, respectively. When measured

immediately before the daily exposure, plasma levels had returned to baseline in females.

3.2.3.1.3 Dermal Exposure

No specific dermal absorption studies have been conducted in rats. However, Kennedy (1985)

treated rats dermally with a total of 10 applications of APFO at doses of 0, 20, 200 or 2,000

mg/kg. Treatment resulted in elevated blood organofluorine levels that increased in a doserelated manner.
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28

3.2.3.2 Serum Pharmacokinetic Parameters in Adult Rats

Serum pharmacokinetic parameters of PFOA have been evaluated in adult Sprague-Dawley rats

following gavage administration, and in Wistar rats following i.v. administration. 

3.2.3.2.1 Oral and Intravenous Exposure in Sprague-Dawley Rats

Kemper (2003) examined the plasma concentration profile of PFOA following gavage and

intravenous administration in sexually mature Sprague-Dawley rats. Male and female rats

(4/sex/group) were administered single doses of PFOA by gavage at dose rates of 0.1, 1, 5, and

25 mg/kg PFOA, and intravenously at a dose rate of 1 mg/kg PFOA. After dosing, plasma was

collected for 22 days in males and 5 days in females. Plasma concentration vs. time data were

then analyzed by non-compartmental pharmacokinetic methods (Tables 2 and 3). Comparison of

AUC for the oral and intravenous 1 mg/kg doses indicated that oral bioavailability of PFOA was

approximately 100%. Plasma elimination curves were linear with respect to time in male rats at

all dose levels, while elimination kinetics were biphasic in females at the 5 mg/kg and 25 mg/kg

dose levels. In males, plasma elimination half-lives were independent of dose level and ranged

from approximately 138 hours to 202 hours. In females, terminal elimination half-lives ranged

from approximately 2.8 hours at the lowest dose to approximately 16 hours at the high dose. To

further characterize plasma elimination kinetics, particularly in male rats, animals were given

oral PFOA at a rate of 0.1 mg/kg, and plasma samples were collected until PFOA concentrations

fell below quantitation limits (24 hours and 2016 hours in females and males, respectively). 

Estimated plasma elimination half-lives in this experiment were approximately 277 hours in

males and 3.4 hours in females (Tables 2 and 3).
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29

Table 2

Pharmacokinetic Parameters in Male Sprague-Dawley Rats 

Following Administration of PFOA by Gavage (Mean (SD))

Parameter 0.1 mg/kg 1 mg/kg 5 mg/kg 25 mg/kg 1 mg/kg 

i.v.

0.1 mg/kg

extended

time

Tmax (hr) 10.25

(6.45)

9.00

(3.83)

15.0

(10.5)

7.5

(6.2)

NA 5.5

(7.0)

Cmax (:g/mL) 0.598

(0.127)

8.431

(1.161)

44.75

(6.14)

160.0

(12.0)

NA 1.08

(0.42)

Lambda z (1/hr) 0.004

(0.001)

0.005

(0.001)

0.0041

(0.0007)

0.0046

(0.0012)

0.004

(0.000)

0.0026

(0.0007)

T1/2 (hr) 201.774

(37.489)

138.343

(31.972)

174.19

(28.92)

157.47

(38.39)

185.584

(19.558)

277.10

(56.62)

AUCINF (hrA:g/mL) 123.224

(35.476)

1194.463

(215.578)

6733.70

(1392.83)

25155.61

(7276.96)

1249.817

(113.167)

206.38

(59.03)

AUCINF/D (hrA:g/mL/mg/kg) 1096.811

(310.491)

1176.009

(206.316)

1221.89

(250.28)

942.65

(284.67)

1123.384

(100.488)

2111.28

(586.77)

Clp (mL/kgAhr) 0.962

(0.240)

0.871

(0.158)

0.85

(0.21)

1.13

(0.31)

0.896

(0.082)

0.51

(0.17)
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30

Table 3

Pharmacokinetic Parameters in Female Sprague-Dawley Rats 

Following Administration of PFOA by Gavage (Mean (SD))

Parameter 0.1 mg/kg 1 mg/kg 5 mg/kg 25 mg/kg

1 mg/kg 

i.v.

0.1 mg/kg

extended

time

Tmax (hr) 0.56

(0.31)

1.13

(0.63)

1.50

(0.58)

1.25

(0.87)

NA 1.25

(0.50)

Cmax (:g/mL) 0.67

(0.07)

4.782

(1.149)

20.36

(1.58)

132.6

(46.0)

NA 0.52

(0.08)

Lambda z (1/hr) 0.231

(0.066)

0.213

(0.053)

0.15

(0.02)

0.059

(0.037)

0.250

(0.047)

0.22

(0.07)

T1/2 (hr) 3.206

(0.905)

3.457

(1.111)

4.60

(0.64)

16.22

(9.90)

2.844

(0.514)

3.44

(1.26)

AUCINF (hrA:g/mL) 3.584

(0.666)

39.072

(10.172)

114.90

(11.23)

795.76

(187.51)

33.998

(7.601)

3.34

(0.32)

AUCINF/D (hrA:g/mL/mg/kg) 31.721

(5.880)

38.635

(10.093)

20.78

(2.01)

29.54

(6.92)

30.747

(6.759)

34.39

(3.29)

Clp (mL/kgAhr) 32.359

(6.025)

27.286

(7.159)

48.48

(4.86)

35.06

(.88)

34.040

(9.230)

29.30

(3.06)

3.2.3.2.2 Intravenous Exposure in Wistar Rats

Kudo et al. (2002) examined the plasma concentration profile of PFOA following i.v.

administration in 9-week old Wistar rats. Male and female rats were injected intraveneously

with 48.63 μmol/kg body weight PFOA, and blood samples were collected at 5 min., 1, 6, 12 hrs,

and 1, 2, 3, 4, 6, 8, 10 and 12 days after injection. A two-compartment open model was used to

estimate serum pharmacokinetic parameters (Table 4). The half-life of PFOA in male rats was

found to be 70 times longer (5.7 days versus 1.9 hours) than in females. In female rats, there

appears to be biphasic elimination of PFOA; the fast phase occurs with a half-life of

approximately 1.9 hours while the slow phase occurs with a half life of approximately 24 hours. 

The total clearance of PFOA in female rats was 44 times higher than that in male rats.

Table 4

Pharmacokinetic Parameters of PFOA in Wistar Rats after i.v. Administration

Parameters Male Female

Total clearance (ml/min per kg) 0.035+0.010 1.551 + 0.559*

T1/2 (day) 5.68+0.99 0.08 + 0.03**

Volume of distribution (ml/kg) 345.6+57.3 211.2+ 28.2*

Values are means + S.D. for three to four rats. Differences were statistically significant between
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31

male and female rats (*, P<0.01: **, P<0.001). 

3.2.3.3 Distribution Studies in Adult Rats

The distribution of PFOA has been examined in tissues of adult rats following administration by

gavage and by i.v. and intraperitoneal (i.p.) injection. PFOA distributes primarily to the liver,

plasma, and kidney, and to a lesser extent, other tissues of the body. It does not partition to the

lipid fraction or adipose tissue. 

3.2.3.3.1 Oral Exposure

Kemper (2003) examined the distribution and clearance of PFOA in tissues of male and female

Sprague-Dawley rats following administration by gavage. Rats were administered 1, 5, and 25

mg/kg 14C-PFOA by oral gavage. Tissue concentrations were determined at the time of

maximum plasma concentration (Tmax) and at the time plasma concentration had fallen to one

half the maximum (Tmax/2). Values for Tmax and Tmax/2 for male and female rats were determined

from pharmacokinetic experiments. Tmax/2 was calculated as Tmax + T1/2. For cases in which

biphasic elimination was evident, the rapid phase T1/2 was used for calculation of Tmax/2. Tissues

from male rats were collected at 10.5 hours (Tmax) and 171 hours (Tmax/2) after dosing. Tissues

from female rats were collected at 1.25 hours (Tmax) and 4 hours (Tmax/2) after dosing. The results

are summarized in Tables 5 and 6 for males and females, respectively. Liver, kidney and blood

were the primary tissues for distribution of 14C-PFOA. In males, the fraction of the dose found

in liver increased from Tmax to Tmax/2, but remained constant or decreased in other tissues. In

females, the fraction of the dose present in all tissues remained constant or decreased between

Tmax and Tmax/2. Liver-to-blood concentration ratios for 14C at Tmax in males were greater than 1,

and increased between Tmax and Tmax/2. Kidney-to-blood concentration ratios at Tmax in females

were approximately 2 at all dose levels and remained relatively constant between Tmax and Tmax/2.
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32

Table 5

Distribution of PFOA in Male Sprague-Dawley Rats after Oral Exposure

(Percent of dose recovered at Tmax and Tmax/2 in tissues)

1 mg/kg 1 mg/kg 5 mg/kg 5 mg/kg 25 mg/kg 25 mg/kg

Tmax Tmax/2 Tmax Tmax/2 Tmax Tmax/2

prostate 0.083±0.039 0.030±0.002 0.071±0.045 0.057±0.020 0.067±0.018 0.028±0.012

skina 14.77±2.135 6.061±0.274 15.565±0.899 7.233±0.430 13.836±0.969 5.419±0.237

blooda 22.148±0.692 8.232±1.218 24.919±1.942 11.140±0.624 22.905±1.177 7.904±1.032

brain 0.071±0.018 0.022±0.002 0.051±0.021 0.023±0.008 0.063±0.007 0.019±0.002

fata 2.281±0.467 0.593±0.136 2.815±0.225 0.916±0.205 2.153±0.430 0.628±0.110

heart 0.451±0.119 0.195±0.024 0.443±0.037 0.252±0.030 0.461±0.053 0.164±0.032

lungs 0.74±0.147 0.341±0.043 0.593±0.376 0.344±0.194 0.863±0.103 0.303±0.057

spleen 0.086±0.011 0.045±0.006 0.096±0.017 0.060±0.007 0.106±0.015 0.042±0.005

liver 21.708±5.627 32.627±3.601 18.750±2.434 25.231±1.289 17.528±0.900 20.145±3.098

kidney 1.949±0.402 1.14±0.215 2.170±0.354 1.212±0.115 2.293±0.286 1.003±0.122

G.I.

tract

2.930±0.929 0.980±0.300 2.508±0.713 1.052±0.202 2.784±0.608 0.808±0.189

G.I.

contents

2.083±0.625 0.239±0.025 2.632±0.934 0.270±0.028 4.186±1.349 0.210±0.084

thyroid 0.008±0.005 0.004±0.003 0.011±0.006 0.004±0.002 0.009±0.002 0.005±0.001

thymus 0.085±0.008 0.051±0.018 0.085±0.012 0.053±0.003 0.120±0.025 0.045±0.010

testes 0.755±0.079 0.356±0.037 0.693±0.180 0.372±0.062 0.623±0.098 0.224±0.031

adrenals 0.019±0.004 0.010±0.001 0.022±0.004 0.009±0.001 0.026±0.004 0.009±0.003

musclea 12.025±0.648 4.984±0.745 13.565±0.576 6.429±0.648 12.855±0.841 4.253±0.358

bonea 3.273±0.538 1.120±0.094 3.047±0.544 1.375±0.169 3.062±0.438 0.906±0.100

Totalb 85.465±6.426 57.026±3.379 88.033±1.420 56.031±1.025 83.937±3.680 42.112±4.740

a

 Percent recovery scaled to whole animal assuming the following: skin=19%, 

 whole blood=7.4%, fat=7%, muscle=40.4%, bone=7.3% of body weight.

b

 Totals are calculated from individual animal data.
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33

Table 6

Distribution of PFOA in Female Sprague-Dawley Rats after Oral Exposure

 (Percent of dose recovered at Tmax and Tmax/2 in tissues)

Tissue

1 mg/kg 5 mg/kg 25 mg/kg

Tmax Tmax/2 Tmax Tmax/2 Tmax Tmax/2

skina 0.434±0.162 0.403±0.096 0.624±0.142 0.307±0.121 0.380±0.166 0.415±0.175

blooda 5.740±1.507 4.438±1.625 8.089±2.080 5.411±1.466 7.158±2.232 6.407±1.406

brain 0.037±0.009 0.047±0.008 0.066±0.019 0.045±0.010 0.058±0.008 0.058±0.018

fata 0.134±0.032 0.164±0.079 0.220±0.111 0.110±0.069 0.147±0.053 0.148±0.065

heart 0.198±0.079 0.253±0.055 0.388±0.057 0.236±0.051 0.317±0.035 0.287±0.069

lungs 0.454±0.148 0.546±0.082 0.827±0.102 0.570±0.179 0.678±0.067 0.775±0.204

spleen 0.063±0.027 0.058±0.006 0.101±0.021 0.060±0.012 0.091±0.007 0.070±0.002

liver 7.060±1.266 6.817±1.537 11.190±2.192 7.176±0.982 10.538±1.723 9.080±0.895

kidney 3.288±0.948 2.769±0.784 4.293±0.771 2.685±0.736 5.867±0.946 4.749±0.393

G.I.

tract

10.699±9.066 8.462±6.519 7.142±2.594 8.255±8.967 6.923±1.846 3.547±1.306

G.I.

contents

21.956±13.48 3.891±2.395 2.896±2.305 5.601±6.165 2.491±1.548 1.121±1.010

thyroid 0.010±0.003 0.016±0.021 0.008±0.002 0.006±0.002 0.009±0.003 0.007±0.002

thymus 0.052±0.017 0.058±0.024 0.105±0.030 0.068±0.021 0.091±0.032 0.077±0.020

ovaries 0.047±0.019 0.048±0.006 0.071±0.012 0.041±0.012 0.071±0.012 0.070±0.012

adrenals 0.014±0.005 0.018±0.004 0.026±0.005 0.015±0.004 0.031±0.005 0.021±0.001

musclea 0.170±0.051 0.258±0.089 0.325±0.010 0.229±0.031 0.441±0.116 0.304±0.099

uterus 0.243±0.091 0.374±0.247 0.354±0.046 0.247±0.068 0.358±0.124 0.365±0.029

bonea 0.101±0.017 0.153±0.052 0.174±0.057 0.142±0.078 0.157±0.072 0.181±0.090

Totalb 50.698±16.48 28.772±10.98 36.897±3.187 31.201±12.63 35.803±2.554 27.680±2.569

a

 Percent recovery scaled to whole animal assuming the following: skin=19%, whole

blood=7.4%, fat=7%, muscle=40.4%, bone=7.3% of body weight.

b

 Totals are calculated from individual animal data.
SAB Review Draft; Do Not Cite or Quote

34

3.2.3.3.2 Intravenous Exposure

Gibson and Johnson (1980) examined the tissue distribution in 10-week old male CD rats

following a single i.v. dose of 13.1 mg/kg 14C-PFOA. The animals were sacrificed 36 days after

dosing. The results are presented in Table 7. PFOA distributed mainly to the liver, plasma and

kidney. 

Table 7

Distribution of PFOA in Male CD Rats After i.v. Exposure

Tissue Mean + SDa

Liver 7.97 + 4.02

Plasma 3.19 + 1.72

Kidney 2.33 + 1.13

Lung 0.79 + 0.51

Red Blood Cells 0.77 + 0.45

Skin 0.40 + 0.28

Spleen 0.33 + 0.19

Bone marrow 0.33 + 0.26

Subcutaneous Fat 0.28 + 0.18

Muscle 0.20 + 0.09

Brain 0.11 + 0.04

Abdominal Fat < 0.12 

a - Total 14C concentration expressed as :g equivalents of 14C-PFOA. 

3.2.3.3.3 Intraperitoneal Exposure

Ylinen et al. (1990) studied the distribution and elimination of PFOA in a limited sample of

tissues following a single i.p. dose of 50 mg/kg in 10 week old Wistar rats (20/sex). Samples

were collected from 2 animals per sex for analysis of PFOA at 12, 24 - 168 (in 24 hour

intervals), 224 and 336 hours after dosing. The concentration of PFOA in the serum and tissues

was determined with capillary gas chromatography equipped with a flame ionization detector

(FID). A mass spectrometer was used in the selected ion monitoring mode when the PFOA

concentration was below the quantitation limit of the FID (1 :g/ml). The half-life of PFOA was

estimated from the linear regression of time and concentration of PFOA in a semilogarithmic

plot. No PFOA was detected in the adipose tissue. In both sexes at 12 hours after

administration, the highest concentration of PFOA was found in the serum, followed by the liver,

kidney, spleen and brain. In the females, the concentration of PFOA in the serum, liver and

kidney decreased in a discontinuous fashion indicating distinct phases. The half-life of PFOA in
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35

the serum, liver, kidney and spleen is presented in Table 8; the values are generally much lower

in females than in males. 

Table 8

Mean Half-life (hours) of PFOA in Wistar Rats After i.p. Exposure

Tissue Males Females

Serum 105 24 

Liver 210 60

Kidney 130 145

Spleen 170 73

Vanden Heuvel et al. (1991b) examined the distribution of PFOA in the serum and a limited

number of tissues following administration of 9.4 :mol/kg 14C-PFOA by i.p. injection to 6-week

old male and female Harlan Sprague-Dawley rats. At various time-points for 28 days following

treatment, four rats per sex were sacrificed; blood was collected by cardiac puncture, and tissues

were removed and frozen. 14C-PFOA-derived radioactivity was quantitated using a Liquid

Scintillation Analyzer. Assuming first order kinetics, tissue elimination rates were calculated as

T1⁄2 = ln2/ke. The distribution and elimination of PFOA-derived 14C in selected tissues is

summarized in Table 9. In the male rats, 21% of the administered dose was present in the liver,

and the next highest concentrations were found in the plasma and kidney. Far lower PFOA

concentrations were found in the heart, testis, fat, and gastrocnemius muscle. In females, the

highest concentrations of PFOA were found in the plasma followed by the kidney, liver and

ovaries. In males, PFOA was eliminated from the liver at a slower rate than the other tissues;

the T1⁄2 for liver was 11 days compared to 8-9 days in most extrahepatic tissues. The rates of

elimination were much faster in the female rats than in the male rats. 

The high concentration of PFOA in the male liver was further examined using a liver perfusion

technique. Liver was infused with 0.08 :mol 14C-PFOA/min over a 48 min period for a total of

3.84 umol 14C. Approximately 11% of the cumulative dose of 14C-PFOA infused was extracted

by the liver in a first pass. At 2 min, the cumulative percent of PFOA extracted by the liver was

33%; that was substantially greater than the 11% cumulative dose of 14C that was extracted after

48 min indicating that first-pass hepatic uptake of PFOA may be saturable. 
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36

Table 9

Distribution and Elimination of PFOA in Selected Tissues in Harlan Sprague-Dawley Rat

after i.p. Exposurea

Tissue

 % Dosea Elimination Rateb Half-life

Male

(2 hrs)

Female

(2 hrs)

Male

(24 hrs)

Female

(24 hrs)

Male 

(day -1)

Female 

(hour -1)

Male 

(day)

Female 

(hour)

Liver 2.03

(0.10)

1.53

(0.25)

2.08

(0.14)

0.06

(0.02)

0.062

(0.005)

0.185

(0.011)

11.3 3.8

Plasma 1.99

(0.03)

2.39

(0.44)

1.63

(0.11)

0.02 0.077

(0.006)

0.242

(0.007)

9.0 2.9

Kidney 0.95

(0.08)

2.00

(0.38)

0.74

(0.03)

0.06

(0.02)

0.074

(0.005)

0.214

(0.012)

9.4 3.2

Heart 0.42

(0.02)

ND 0.39

(0.03)

ND 0.066

(0.006)

ND 10.4 ND

Gastrocnemius 0.26

(0.01)

ND 0.21

(0.01)

ND 0.076

(0.007)

ND 9.2 ND

Fatc 0.32

(0.02)

ND 0.27

(0.04)

ND 0.087

(0.008)

ND 8.0 ND

Testis/

Ovary

0.33

(0.01)

0.53

(0.10)

0.27

(0.03)

0.05

(0.01)

0.078

(0.006)

ND 9.0 ND

ND - Not determined

a- Values are percent dose of PFOA-derived 14C per gram tissue; mean (SEM); n=4; values with

no SEM indicate that SEM < 0.01

b- Mean (SEM); n=32 for males and n=16 for females

c-Epididymal fat pad

3.2.3.4 Metabolism Studies in Adult Rats

Several studies have examined metabolism of PFOA. However, no studies show clear evidence

of metabolism. Ophaug and Singer (1980) found no change in ionic fluoride level in the serum

or urine following oral administration of PFOA to female Holtzman rats. Ylinen et al. (1989)

found no evidence of phase II metabolism of PFOA following a single intraperitoneal PFOA

dose (50 mg/kg) in male and female Wistar rats.

Vanden Heuvel et al. (1991b) investigated the metabolism of PFOA in Harlan Sprague-Dawley

rats administered 14C-PFOA (9.4 :mol/kg, i.p.). Pooled daily urine samples (0-4 days posttreatment) and bile extracts analyzed by HPLC contained a single radioactive peak eluting

identically to the parent compound. Tissues were taken from rats treated 4, 14, and 28 days after

treatment to determine the presence of PFOA-containing lipid conjugates. Only the parent

compound was present in rat tissues; no PFOA-containing hybrid lipids were detected. Fluoride

concentrations in plasma and urine before and after PFOA treatment were unchanged, indicating

that PFOA does not undergo defluorination.
SAB Review Draft; Do Not Cite or Quote

37

3.2.3.5 Elimination Studies in Adult Rats

In adult rats, there is evidence of enterohepatic circulation of PFOA. The urine is the major

route of excretion of PFOA in the female rat, while the urine and feces are both main routes of

excretion in male rats. There are gender differences in the elimination of PFOA in rats. The

rapid excretion of PFOA by female rats is due to active renal tubular secretion (organic acid

transport system); this renal tubular secretion is believed to be hormonally controlled. Hormonal

changes during pregnancy do not appear to change the rate of elimination in rats.

3.2.3.5.1 Enterohepatic Circulation

Johnson et al. (1984) investigated the effect of feeding cholestyramine to rats on the elimination

of APFO. Since APFO exists as an anion at physiologic pH, it would be expected to complex

with cholestyramine. Ten male Charles River CD rats, 12 weeks of age, were given a single i.v.

injection of 13 mg/kg 14C-APFO. Five rats were given 4% cholestyramine in feed. Urine and

feces samples were collected at intervals for 14 days, at which time the animals were sacrificed

and liver samples were collected. At 14 days post dose, the mean percentage of PFOA

eliminated in the feces of cholestyramine-treated rats was 9.8-fold the mean percentage

eliminated in the feces of rats that did not receive cholestyramine. Excretion in urine was 41%

of the administered dose for cholestyramine treated rats and 67% for rats that did not receive

cholestyramine. 14C in the liver equaled 4% or 12.1 ± 2.1 :g eq/g in cholestyramine treated rats

and 8 % or 22.3 ± 6.2 :g eq/g in rats that did not receive cholestryamine. In plasma, the levels

were 5.1 ± 1.7 :g eq/ml in cholestyramine treated rats and 14.7±6.8 :g eq/ml in rats that did not

receive cholestyramine. In red blood cells, the levels were 1.8 ± 0.7 :g eq/ml in cholestryamine

treated rats and 4.2 ± 2.4 :g eq/ml in rats that did receive cholestryamine. The high

concentration of 14C-APFO in the liver at 2 weeks after dosing and the fact that cholestyramine

treatment enhances fecal elimination of 14C nearly 10-fold suggests that there is enterohepatic

circulation of PFOA.

3.2.3.5.2 General Elimination Studies

3.2.3.5.2.1 Oral Exposure

Kemper (2003) investigated the elimination of PFOA in male and female Sprague-Dawley rats

(4 rats/sex/group) administered a single dose of 14C-PFOA by oral gavage at dose levels of 1, 5,

and 25 mg/kg. Urine and feces were collected for 28 days in males and 7 days in females. Urine

was the primary route of excretion of 14C in both sexes, accounting for 43-62% of the

administered dose in males and 76-84% of the administered dose in females. Cumulative

recovery of 14C in the urine increased gradually over the 28 days in male rats, but was essentially

complete in female rats within the first 72 hours. Fecal excretion of 14C accounted for 6-14% of

the dose in males and 2-6% of the dose in females. Pilot experiments demonstrated that 14C was

not eliminated as either 14CO2 or volatile organic compounds in 14C-PFOA-treated rats. 

Pretreatment of rats with 1 mg/kg-day PFOA for 14 days had little or no effect on the excretion

of a challenge dose of 1 mg/kg 14C-PFOA.

3.2.3.5.2.2 Intravenous Exposure

Gibson and Johnson (1980) examined the excretion of total 14C in male and female CD rats after

a single i.v. dose of 14C-PFOA. The mean dose for females was 16.7 mg/kg while that for males

was 13.1 mg/kg. Female rats excreted essentially all of the administered dose via the urine in the
SAB Review Draft; Do Not Cite or Quote

38

24 hours after treatment. During the same time period, male rats excreted only 20% of the total

dose. Male rats excreted 83% of the total dose via the urine and 5.4% via the feces by 36 days

post dose. No radioactivity was detected in tissues of female rats at 17 days post dose; 2.8% of

the total dose was detected in the liver of male rats and 1.1% in the plasma at 36 days post dose

with lower levels equaling < 0.5% of the total dose in other organs.

3.2.3.5.3 Elimination Studies in the Pregnant Rat

Hormonal changes during pregnancy do not appear to cause a change in the rate of elimination

of 14C after oral administration of a single dose of 14C-APFO (Gibson and Johnson, 1983). At 8

or 9 days after conception, four pregnant CD rats and two nonpregnant female CD rats were

given a mean dose of 15 mg/kg 14C-APFO. Individual urine samples were collected at 12, 24,

36, and 48 hours post dose and analyzed for 14C content. Essentially all of the 14C was

eliminated via the urine within 24 hours for both groups of rats.

3.2.3.5.4 Studies on the Mechanism of the Gender Difference in Elimination in Adult Rats

Several studies have been conducted to elucidate the cause of the gender difference in rats in the

elimination of PFOA. Hanhijarvi et al. (1982) conducted a series of studies to examine the

effect of probenecid, which inhibits the renal active secretion system for organic acids, on the

elimination of PFOA in male and female Holtzman rats. In the first study, 4 male and 6 female

Holtzman rats were administered 2 mg of nonionic fluorine as PFOA by gavage. Seven female

rats were administered 2 ml distilled water as controls. The animals were then placed in

metabolism cages and urine was collected until the animals were sacrificed at 24 hr by cardiac

puncture. Serum was collected. Ionic fluoride and total fluorine content of serum and urine

were determined, and nonionic fluorine was calculated as the difference. Twenty-four hours

after oral administration of PFOA, female rats had excreted 76 ± 2.7% of the dose in the urine

and had a mean serum nonionic fluorine level of 0.35 ± 0.11 :g/ml, while male rats had excreted

only 9.2 ± 3.5% of the dose and had a mean serum nonionic fluorine level of 44.0 ± 1.7 :g/ml. 

A mean of 97.5±0.25% of the PFOA was bound in the plasma of both male and female rats. 

In the second study, Hanhijarvi et al. (1982) examined the effect of probenecid on the clearance

of PFOA and inulin. Holtzman rats were anesthetized and the femoral artery was cannulated for

continuous infusion of 5% mannitol in isotonic saline, while the femoral artery was cannulated

for drawing blood samples. The urinary bladder was also cannulated for serial collections of

urine. When the urine and serum collections for the clearance study were complete, 65-68

mg/kg probenecid was administered by i.p. injection and after 20 - 30 minutes, additional 10

minute clearance tests were performed. Administration of probenecid reduced the PFOA/inulin

clearance ratio in females from 14.5 to 0.46. PFOA clearance was reduced from 5.8 to 0.11

ml/min/100g. Net PFOA excretion was reduced from 4.6 :g/min/100g to 0.13 :g/min/100g. In

male rats, however, the PFOA/inulin clearance ratio and the net excretion of PFOA were

virtually unaffected by probenecid. In the males, PFOA clearance was 0.17 ml/min/100g, the

PFOA/inulin clearance ratio was 0.22, and net PFOA excretion was 0.17 :g/min/mg.

Finally, Hanhijarvi et al. (1982) examined the cumulative excretion of PFOA over a 7-hour

period. Holtzman rats were dosed i.v. with a mixture of 10%-20% radiolabeled-PFOA and 80-

90% unlabeled PFOA. Mannitol (5%) was infused and urine specimens were collected over 30-

min intervals. The effect of probenecid was assessed by administering 65-68 mg/kg by i.p.

injection at least 30 min prior to the administration of PFOA. Female rats excreted 76% of the

administered dose of PFOA, while males excreted only 7.8% of the administered dose over a 7-
SAB Review Draft; Do Not Cite or Quote

39

hr period. Probenecid administration modified the cumulative excretion curve for males only

slightly. However, in females probenecid markedly reduced PFOA elimination to 11.8%. The

authors concluded that the female rat possesses an active secretory mechanism which rapidly

eliminates PFOA from the body. 

Ylinen et al. (1989) studied the urinary excretion of PFOA in male Wistar rats after castration

and estradiol administration. Twenty male rats were castrated at 28 days of age and were used in

tests of PFOA excretion 5 weeks later. Ten castrated and 10 intact males were given 500 :g/kg

estradiol valerate by s.c. injection every second day for 14 days before administration of PFOA.

PFOA was administered as a single i.p. injection at 50 mg/kg. Urine was collected in metabolism

cages for 96 hr after PFOA administration. Blood samples were collected by cardiac puncture.

Six female rats were also included in the experiment. Castration and administration of estradiol

to the male rats had a significant stimulatory effect on the urinary excretion of PFOA. During the

first 24 hours, female rats excreted 72 ± 5% of the administered dose of PFOA, whereas the

intact males excreted only 9 ± 4%. After the estradiol treatment, both the intact and castrated

males excreted PFOA in amounts similar to females, 61 ± 19% and 68 ± 14%, respectively. The

castrated males without estradiol treatment excreted 50 ± 13% of the administered dose of PFOA

in the urine. This was faster than the intact males but less than the females and the estrogen

treated males. At the end of the test, the concentration of PFOA in the serum of intact males was

17- 40 times higher than the concentration PFOA in the serum of other groups. There were no

statistically significant differences in the serum concentrations between the other groups. PFOA

was similarly bound by the proteins in the serum of males and females.

Vanden Heuvel et al. (1992a) investigated whether androgens or estrogens are involved in the

marked sex-differences in the urinary excretion of PFOA. Castrated Harlan Sprague-Dawley

male rats were given 9.4 :mol/kg 14C-PFOA by i.p. injection. Castration increased the

elimination of PFOA in the urine (36% of the dose was eliminated in 4 days versus 16% in

controls), suggesting that a factor produced by the testis is responsible for the slow elimination

of PFOA in male rats. Castration plus 17B-estradiol had no further effect on PFOA elimination

whereas castration plus testosterone replacement at the physiological level reduced PFOA

elimination to the same level as rats with intact testis. Thus, in male rats, testosterone exerts an

inhibitory effect on renal excretion of PFOA. In female rats, neither ovariectomy or

ovariectomy plus testosterone affected the urinary excretion of PFOA, demonstrating that the

inhibitory effect of testosterone on PFOA renal excretion is a male-specific response. 

Probenecid, which inhibits the renal transport system, decreased the high rate of PFOA renal

excretion in castrated males but had no effect on male rats with intact testis.

Kudo et al. (2002) examined the role of sex hormones on the renal clearance (CLR) of PFOA and

the renal mRNA levels of specific organic anion transporters in male and female Wistar rats. 

Castration of male rats caused a 14-fold increase in CLR of PFOA. The elevated PFOA CLR in

castrated males was reduced by treating them with testosterone. Treatment of male rats with

estradiol increased the CLR of PFOA. In female rats, ovariectomy caused a significant increase

in CLR of PFOA, which was reduced by estradiol treatment. Treatments of female rats with

testosterone reduced the CLR of PFOA. Treatment with probenecid, a known inhibitor of

organic anion transporters, markedly reduced the CLR of PFOA in male rats, castrated male rats,

and female rats. To identify the transporter molecules that are responsible for PFOA transport in

the rat kidney, renal mRNA levels of specific organic anion transporters were determined in

male and female rats under various hormonal states and compared with the CLR of PFOA. The

level of OAT2 mRNA in male rats was only 13% that in female rats. Castration or estradiol

treatment increased the level of OAT2 mRNA whereas treatment of castrated male rats with
SAB Review Draft; Do Not Cite or Quote

40

testosterone reduced it. Ovariectomy of female rats significantly increased the level of OAT3

mRNA. Multiple regression analysis of the data suggested that organic anion transporter 2

(OAT2) and OAT3 are responsible for urinary elimination of PFOA in the rat. 

3.2.4 Metabolism and Pharmacokinetic Studies in Immature Rats

No studies have been conducted to specifically examine the absorption, metabolism or

elimination of PFOA in the developing rat. However, recent studies have been conducted to

examine the concentrations of PFOA in the developing Sprague-Dawley rat, and to determine

when the gender difference in elimination of PFOA becomes apparent. In addition, several

studies have examined the serum and tissue distribution of PFOA in newly weaned Wistar rats. 

These studies have shown that PFOA readily crosses the placenta and is present in the breast

milk of rats. During lactation and immediately after weaning, the elimination of PFOA is similar

in males and females. In the male rat between 4-5 weeks of age, the factor(s) responsible for the

gender difference develop, and the rats assume the adult male elimination profile. In addition,

distribution studies in the postweaning rat have shown that PFOA is distributed primarily to the

serum, liver, and kidney.

3.2.4.1 PFOA Levels During Pregnancy and Lactation

Mylchreest (2003) examined PFOA levels during gestation and lactation. Pregnant SpragueDawley rats were dosed with 0, 3, 10 or 30 mg/kg-day APFO during days 4-10, 4-15, or 4-21 of

gestation, or from gestation day 4 to lactation day 21. Clinical observations and body weights

were recorded daily. On gestation days 10, 15, and 21, 5 rats/group/time point were sacrificed

and the number, location and type of implantation sites were recorded. Embryos were collected

on day 10, and placentas, amniotic fluid, and embryos/fetuses were collected on days 15 and 21. 

Maternal blood samples were collected at 2 hours + 30 minutes post-dose. The remaining 5 rats

per group were allowed to deliver. On lactation days 0, 3, 7, 14, and 21, the pups were counted,

weighed (sexes separate), and examined for abnormal appearance and behavior. Randomly

selected pups were sacrificed and blood samples were collected. On lactation days 3, 7, 14, and

21, the dams were anesthetized and milk and blood samples were collected; dams were removed

from their litters 1-2 hours prior to collection. Plasma, milk, amniotic fluid extract, and tissue

homogenates (placenta, embryo, and fetus) supernatants were analyzed for PFOA concentrations

by HPLC-MS.

All dams survived and there were no clinical signs of toxicity. In the 30 mg/kg-day group, mean

body weight gain was approximately 10% lower than the control group during gestation, and

mean body weights were approximately 4% lower than controls throughout gestation and

lactation. The number of implantation sites, resorptions, and live fetuses were comparable

among groups on days 10, 15, and 21 of gestation. One dam in the 3 mg/kg-day group and two

dams in the 30 mg/kg-day group delivered small litters (litter size of 3-6 pups as compared to 12-

19 pups/litter in the control group); however, given the small sample size the biological

significance of this finding is unclear. There were no clinical signs of toxicity in the pups, and

pup survival and pup body weights were comparable among groups.

Maternal PFOA levels during gestation and lactation are presented in Table 10. Maternal plasma

levels at 2 hrs post-dosing (approximately the time of peak blood levels following a gavage

dose) were fairly similar during the course of the study with a mean level of 11.2, 26.8, and 66.6

:g/ml in the 3, 10, and 30 mg/kg-day groups, respectively; PFOA levels in the control group

were below the limit of quantitation (0.05 :g/ml). The concentration of PFOA in the milk was
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41

also fairly similar throughout lactation and was approximately 1/10th of the PFOA levels in the

plasma; the mean values were 1.1, 2.8, and 6.2 :g/ml in the 3, 10, and 30 mg/kg-day groups,

respectively. 

Table 10

Maternal PFOA Levels (:g/ml) During Gestation and Lactationa

Exposure Period Sample Time 3 mg/kg-day 10 mg/kg-day 30 mg/kg-day

GD 4 - GD 10 GD 10 plasma 8.53 + 1.06 23.32 + 2.15 70.49 + 8.94

GD 4 - GD 15 GD 15 plasma 15.92 + 12.96 29.40 + 14.19 79.55 + 3.11

GD 4 - GD 21 GD 21 plasma 14.04 + 2.27 34.20 + 6.68 76.36 + 14.76

GD 4 - LD 3 LD 3 - plasma

 - milk

11.01 + 2.11

 1.07 + 0.26

22.47 + 2.74

 2.03 + 0.33

54.39 + 17.86

 4.97 + 1.20

GD 4 - LD 7 LD 7 - plasma

 - milk

10.09 + 2.90

 0.94 + 0.22

25.83 + 2.07

 2.74 + 0.91

66.91 + 11.82

 5.76 +1.26

GD 4 - LD 14 LD 14 - plasma

 - milk

 9.69 + 0.92

 1.15 + 0.06

23.79 + 2,81

 3.45 + 1.18

54.65 + 11.63

 6.45 + 1.38

GD 4 - LD 21 LD 21 - plasma

 - milk

 9.04 + 1.01

 1.13 + 0.08

28.84 + 5.15

 3.07 + 0.51

64.13 + 1.45

 7.48 + 1.63

NA Average plasma

Average milk

11.19 + 2.76

 1.07 + 0.09

26.84 + 4.21

 2.82 + 0.60

66.64 + 9.80

 6.16 + 1.06

a- mean + SD; samples were from 5 dams/group/time point and were collected 2 hrs post-dosing

PFOA levels in the placenta, amniotic fluid, embryo, fetus, and pup plasma are presented in

Table 11. The levels of PFOA in the placenta on gestation day 21 were approximately twice the

levels observed on gestation day 15, and the levels of PFOA in the amniotic fluid were

approximately four times higher on day 21 than on day 15. The concentration of PFOA in the

embryo/fetus was highest in the day 10 embryo and lowest in the day 15 embryo; PFOA levels

in the day 21 fetus were intermediate. The concentration of PFOA in the plasma of the day 21

fetus were approximately half the levels observed in the maternal plasma; the mean values were

5.9, 14.5, and 33.1 :g/ml in the 3, 10, and 30 mg/kg-day groups, respectively. Pup plasma levels

decreased until lactation day 7, and were thereafter similar to the levels observed in the milk.
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42

Table 11

PFOA Concentrations (:g/ml) During Gestation and Lactation in Sprague-Dawley Ratsa

Exposure Period Tissue 3 mg/kg-day 10 mg/kg-day 30 mg/kg-day

GD 4 - GD 10 GD 10 - embryo 1.40 + 0.30 3.33 + 0.81 12.49 + 3.50

GD 4 - GD 15 GD 15 - placenta

 - amniotic fluid

 - embryo

2.22 + 1.79

0.60 + 0.69

0.24 + 0.19

 5.10 + 1.70

 0.70 + 0.15

 0.53 + 0.18

13.22 + 1.03

 1.70 + 0.91

 1.24 + 0.22

GD 4 - GD 21 GD 21 - placenta

 - amniotic fluid

 - fetus

 - fetal plasma

3.55 + 0.57

1.50 + 0.32

1.27 + 0.26

5.88 + 0.69

 9.37 + 1.76

 3.76 + 0.81

 2.61 + 0.37

14.48 + 1.51

24.37 + 4.13

 8.13 + 0.86

 8.77 + 2.36

33.11 + 4.64

GD 4 - LD 3 LD 3 - pup plasma 2.89 + 0.70 5.94 + 1.44 11.96 + 1.66

GD 4 - LD 7 LD 7 - pup plasma 0.65 + 0.20 2.77 + 0.58 4.92 + 1.28

GD 4 - LD 14 LD 14 - pup plasma 0.77 + 0.10 2.22 + 0.38 4.91 + 1.12

GD 4 - LD 21 LD 21 - pup plasma 1.28 + 0.72 3.25 + 0.52 7.36 + 2.17

a - mean + SD; samples were pooled by litter and were collected 2 hrs post-dosing

3.2.4.2 PFOA Levels in the Postweaning Rat

Han (2003) examined the relationship between age and plasma PFOA concentrations in the

postweaning Sprague-Dawley rat. Four to eight week old rats (10/sex/time period) were

administered a single dose of 10 mg/kg-day APFO by gavage. Blood samples were collected 24

hours after dosing and the plasma concentration of PFOA was measured by HPLC-MS. At four

weeks of age, the concentration of plasma PFOA was approximately 2.7 times higher in males

than in the females (Table 12). In females, the concentration of plasma PFOA decreased by 2.7

fold between 4 and 5 weeks of age, and thereafter remained fairly steady. In males, the

concentration of plasma PFOA increased by 5.4 fold between 4 and 5 weeks of age, and

thereafter remained fairly steady. Between 5 and 8 weeks of age, the PFOA plasma

concentrations were 34.7 - 65.1 fold higher in males than in females of the same age. Thus, it

appears that the elimination of PFOA is similar among males and females until week 4, and

between weeks 4 and 5 the maturation of the factor(s) responsible for the gender difference in

elimination of PFOA occurs in the male rat. 
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Table 12

Plasma PFOA Concentrations (:g/ml) in Postweaning Sprague-Dawley Ratsa

Age (weeks) Males Females

4 7.32 + 1.01 2.68 + 0.64

5 39.24 + 3.89 1.13 + 0.46

6 43.19 + 3.79 1.18 + 0.52

7 37.12 + 4.07 0.57 + 0.29

8 38.55 + 5.44 0.81 + 0.27

a - mean + SD; samples from 10 animals/sex/group

3.2.4.3 Serum and Tissue Distribution in Immature Wistar Rats Following Oral Exposure

Ylinen et al. (1990) administered newly weaned Wistar rats (18/sex/group) doses of 3, 10, and

30 mg/kg-day PFOA by gavage for 28 days. At necropsy, serum was collected as well as the

brain, liver, kidney, lung, spleen, ovary, testis, and adipose tissue. The concentration of PFOA

in the serum and tissues was determined with capillary gas chromatography equipped with a

flame ionization detector (FID). A mass spectrometer was used in the selected ion monitoring

mode when the PFOA concentration was below the quantitation limit of the FID (1 :g/ml). The

concentration of PFOA in the serum and tissues following 28 days of administration is presented

in Table 13. PFOA was not detected in the adipose tissue. The concentrations of PFOA in the

serum and tissues were much higher in males than in females. In the males, the levels of PFOA

in the serum and tissues were generally lower in the 30 mg/kg-day group than in the 10 mg/kgday group due to increased urinary elimination in the 30 mg/kg-day group.
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Table 13

Tissue Distribution of PFOA in Wistar Rats after 28 Days of Treatment

Tissue

 Malesa Femalesa

 3 

mg/kg-day

 10 

mg/kg-day

 30 

mg/kg-day

 3 

mg/kg-day

 10 

mg/kg-day

 30

mg/kg-day

Serum 48.6+10.3 87.27+20.09 51.65+1.47 2.4b 12.47+4.07 13.92+6.06

Liver 39.9+7.25 51.71+11.18 49.77+10.76 1.81+0.49 3.45+1.36 6.64+2.64

Kidney 1.55+0.71 40.56+14.94 39.81+17.67 0.06+0.02 7.36+3.19 12.54+8.24

Spleen 4.75+1.66 7.59+3.5 4.1+1.57 0.15+0.04 0.38+0.17 1.59+0.49

Lung 2.95+0.54 22.58+4.59 23.71+5.42 0.24b 0.22+0.15 0.75+0.26

Brain 0.398+0.144 1.464+0.211 0.71+0.32 < LOQc 0.029+0.019 0.044+0.018

Ovary < LOQ 0.41+0.27 1.16+0.58

Testis 6.24 + 2.04 9.35 + 4.02 7.22 + 3.17

a- N = 6, Mean + SD, :g/ml

b - N=3, no SD

c - below the limit of quantitation

Hanhijarvi et al. (1987) administered PFOA by gavage to 48 newly-weaned Wistar rats

(6/sex/group) at 0, 3, 10, and 30 mg/kg-day for 28 consecutive days and determined the serum

levels of PFOA. At the end of the study, blood was collected via cardiac puncture; PFOA levels

were determined by gas chromatography. At each dose level, the mean PFOA concentrations in

the plasma of the male rats were significantly higher than those of the female rats. The mean

plasma PFOA concentrations for the male rats were 48.6 ± 26.5 :g/ml, 83.1 ± 24.7 :g/ml, and

53.4 ± 11.2 :g/ml, respectively for the 3, 10 and 30 mg/kg-day dose levels. The corresponding

figures for female rats were 2.43 ± 5.96 :g/ml, 11.3 ± 8.59 :g/ml, and 9.06 ± 8.80 :g/ml

respectively, for the 3, 10 and 30 mg/kg-day dose levels. 

3.2.5 Comparative Studies of Protein Binding in Humans, Non-Human Primates, and Rats

It has been suggested that PFOA circulates in the body by noncovalently binding to plasma

proteins. Several studies have investigated the binding of PFOA to plasma proteins of rats,

humans or monkeys to gain understanding of its absorption, distribution and elimination, and

species and gender differences.

Protein binding of plasma from cynomolgus monkeys, rats, and humans was tested with PFOA

(SRI, 2003). The results are summarized in Table 14. Most PFOA binds to human serum

albumin compared to other protein components of human plasma at physiological

concentrations. Rat, human, and monkey plasma all bind PFOA at 97- 100% at tested

concentrations ranging from 1-500 ppm.
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Table 14 

% Protein Binding to Rat, Human and Monkey Plasma

Nominal Concentration

(ppm)

Rat Monkey Human

1 ~100 ~100 ~100

10 99.5 99.8 99.9

100 98.6 99.8 99.9

250 97.6 99.8 99.6

500 97.3 99.5 99.4

% Binding values reported as “~100" reflect a nonquantifiable amount of test article in the

plasma water BQL<6.25 ng/ml

Han et al. (2003) investigated the binding of PFOA to rat and human plasma proteins in vitro. 

Rats treated in vivo and then sacrificed showed no gender difference in the binding of PFOA to

serum, though the persistence of PFOA in vivo is much greater in male than female rats. The

authors conclude that there is no correlation between the PFOA persistence and binding of the

PFOA to rat serum. The primary PFOA binding protein in plasma was serum albumin.

However, the method used (ligand blotting) would not theoretically allow the identification of

low abundance proteins with high affinity for PFOA. Further investigation of purified rodent

and human serum albumin binding using labeled 19F NMR allowed the calculation of

disassociation constants for PFOA binding to rodent and human serum albumin. No significant

differences between binding to the two proteins was detected (Table 15). 

Table 15

Dissociation Constants (Kd) of Binding between PFOA and Rodent Serum Albumin (RSA) and

Human Serum Albumin (HSA) and the Number of PFOA Binding Sites (n) on RSA and HSA

Parameter Method RSA HSA

Kd (mM) NMRa 0.29 + 0.10c

Kd (mM) micro-SECb 0.36 + 0.08c 0.38 + 0.04

n micro-SECb 7.8 + 1.5 7.2 + 1.3 a

Average of the two Kd values (0.31 + 0.15 and 0.27 + 0.05 mM) obtained by NMR. bValues

were obtained from three independent experiments and their standard deviations are shown. cOn

the basis of the result of unpaired t-test at 95% confidence interval, the difference of Kd values

determined by NMR and micro-SEC is statistically insignificant.

3.2.6 Metabolism and Pharmacokinetic Studies in Other Test Species

There is limited information on the metabolism and pharmacokinetics of PFOA in mice, rabbits

and dogs. No specific pharmacokinetic studies have been conducted in mice. However,

toxicology studies in mice indicate that PFOA is absorbed, and furthermore, there does not
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46

appear to be a gender difference in elimination. For example, Sohlenius et al. (1992) exposed

male and female C57B1/6 mice to dietary levels of 0.02% PFOA for one week. There was a

significant decrease in mean body weight and a significant increase in absolute liver weight; the

response was similar in male and female mice. 

In rabbits, there is no information available on the metabolism or elimination of PFOA by any

route of exposure. No specific studies of the absorption of PFOA have been conducted

following oral or inhalation exposure. However, there is evidence that PFOA is absorbed

following dermal exposure. O’Malley and Ebbins (1981) treated male and female New Zealand

White rabbits dermally with doses of 100, 1,000 and 2,000 mg/kg APFO for 14 days. Mortality

was 100% (4/4) in the 2,000 mg/kg group, 75% (3/4) in the 1,000 mg/kg group and 0% (0/4) in

the 100 mg/kg group. Similarly, Kennedy (1985) treated rabbits dermally with a total of 10

applications of APFO at doses of 0, 20, 200 or 2,000 mg/kg. Treatment resulted in elevated

blood organofluorine levels that increased in a dose-related manner.

In addition, limited information is available on the serum and liver distribution of PFOA in

rabbits following i.v. administration. Johnson (1995a) administered individual female rabbits

intravenous doses of 0, 4, 16, 24 and 40 mg/kg tetrabutyl ammonium salt of PFOA. The animal

given 40 mg/kg died within 5 minutes of treatment. All other animals appeared normal

throughout the study. Serum samples were analyzed for total organic fluorine at 2, 4, 6, 8, 12,

24, and 48 hours post dose. At 2 hrs, serum organic fluorine levels in the rabbits that received 0,

4, 16, and 24 mg/kg were 1.25 :g/ml, 4.09 :g/ml, 14.9 :g/ml, and 41.0 :g/ml, respectively. 

There was a rapid decrease of total organic fluorine in the serum with time; it was non-detectable

at 48 hr. The biological half-life was on the order of 4 hours. The total organic fluorine levels in

whole liver at 48 hr post dose for the rabbits that received 0 mg/kg, 4 mg/kg, 16 mg/kg, and 24

mg/kg were 20 :g, 43 :g, 66 :g, and 54 :g, respectively.

There is no information on the specific absorption, metabolism or distribution of PFOA in dogs

by any route of exposure. One study examined the elimination of PFOA in dogs following i.v.

administration. Hanhijarvi et al. (1988) administered beagle dogs (3/sex) an i.v. injection of 30

mg/kg of PFOA followed by continuous infusion with 5% mannitol. Urine and blood were

collected at 10 minute intervals for 60 min. Probenecid was then administered by i.v. injection,

and urine and blood samples were collected as before. Renal clearance of PFOA was calculated

for the before and after probenecid injection periods. Four additional dogs (2/sex) were given 30

mg/kg of PFOA by i.v. injection. These dogs were kept in metabolism cages, and blood samples

were collected intermittently for 30 days. The renal clearance rate was approximately 0.03

ml/min/kg. Probenecid significantly reduced the PFOA clearance rate in both sexes, indicating

an active secretion mechanism for PFOA. The plasma half-life of PFOA was 473 hr before

probenecid administration and 541 hr after in male dogs, and 202 hr before probenecid and 305

hr after in the female dogs. 

3.3 Acute Toxicity Studies in Animals

Dean and Jessup (1978) reported an oral LD50 of 680 mg/kg and 430 mg/kg for male and female

CD rats, respectively. Glaza (1997) reported an oral LD50 of greater than 500 mg/kg in male

Sprague-Dawley rats and between 250 and 500 mg/kg in females. Gabriel (1976d) reported an

oral LD50 of less than 1000 mg/kg for male and female Sherman-Wistar rats. Rusch (1979)

reported no mortality in male or female Sprague-Dawley rats following inhalation exposure to

18.6 mg/L for one hour. The dermal LD50 in New Zealand White rabbits was determined to be

greater than 2000 mg/kg (Glaza, 1995).
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47

APFO is an ocular irritant in rabbits when the compound is not washed from the eyes (Gabriel,

1976b, 1976e), but is not an irritant in rabbits when washed from the eye (Gabriel, 1976a). 

Markoe (1983) found APFO to be a skin irritant in rabbits, while Gabriel (1976c) did not.

3.4 Mutagenicity Studies

APFO was tested twice (Lawlor, 1995; 1996) for its ability to induce mutation in the Salmonella

– E. coli/mammalian-microsome reverse mutation assay. The tests were performed both with

and without metabolic activation. A single positive response seen at one dose level in S.

typhimurium TA1537 when tested without metabolic activation was not reproducible. APFO

did not induce mutation in either S. typhimurium or E. coli when tested either with or without

mammalian activation. APFO did not induce chromosomal aberrations in human lymphocytes

when tested with and without metabolic activation up to cytotoxic concentrations (Murli, 1996c;

NOTOX, 2000). Sadhu (2002) recently reported that APFO did not induce gene mutation when

tested with or without metabolic activation in the K-1 line of Chinese hamster ovary (CHO) cells

in culture.

Murli (1996b) tested APFO twice for its ability to induce chromosomal aberrations in CHO

cells. In the first assay, APFO induced both chromosomal aberrations and polyploidy in both the

presence and absence of metabolic activation. In the second assay, no significant increases in

chromosomal aberrations were observed without activation. However, when tested with

metabolic activation, APFO induced significant increases in chromosomal aberrations and in

polyploidy (Murli, 1996b). 

 

APFO was tested in a cell transformation and cytotoxicity assay conducted in C3H 10T1⁄2 mouse

embryo fibroblasts. The cell transformation was determined as both colony transformation and

foci transformation potential. There was no evidence of transformation at any of the dose levels

tested in either the colony or foci assay methods (Garry & Nelson, 1981). 

APFO was tested twice in the mouse micronucleus assay. APFO did not induce any significant

increases in micronuclei and was considered negative under the conditions of this assay (Murli,

1996a). 

3.5 Repeat Dose Studies in Animals

3.5.1 Subchronic Studies in Non-Human Primates

Goldenthal (1978b) administered rhesus monkeys (2/sex/group) doses of 0, 3, 10, 30 or 100

mg/kg-day APFO by gavage for 90 days. Animals were observed twice daily and body weights

were recorded weekly. Blood and urine samples were collected once during the control period,

and at 1 and 3 months of the study for hematology, clinical chemistry and urinalysis. Organs

and tissues from animals that were sacrificed at the end of the study and from animals that died

during the treatment period were weighed, examined for gross pathology and processed for

histopathology. 

All monkeys in the 100 mg/kg-day group died during the study. The first death occurred during

week 2; all animals were dead by week 5. Signs and symptoms which first appeared during

week 1 included anorexia, frothy emesis which was sometimes brown in color, pale face and

gums, swollen face and eyes, slight to severe decreased activity, prostration and body trembling.

Three monkeys from the 30 mg/kg-day group died during the study; one male died during week
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48

7 and the two females died during weeks 12 and 13. Beginning in week 4, all four animals

showed slight to moderate and sometimes-severe decreased activity. One monkey had emesis

and ataxia, swollen face, eyes and vulva, as well as pallor of the face and gums. Beginning in

week 6, two monkeys had black stools and one monkey had slight to moderate dehydration and

ptosis of the eyelids. No monkeys in the 3 or 10 mg/kg-day groups died during the study. One

monkey in the 10 mg/kg-day group was anorexic during week 4, had a pale and swollen face in

week 7 and had black stools for several days in week 12. Animals in the 3 mg/kg-day group

occasionally had soft stools or moderate to marked diarrhea; frothy emesis was also occasionally

noted in this group. 

Changes in body weight were similar to the controls for animals from the 3 and 10 mg/kg-day

groups. Monkeys from the 30 and 100 mg/kg-day groups lost body weight after week 1. At the

end of the study, this loss was statistically significant for the one surviving male in the 30

mg/kg-day group (2.30 kg vs 3.78 kg for the control). The results of the urinalysis, and

hematological and clinical chemistry analyses were comparable for the control and the 3 and 10

mg/kg-day groups at one and three months. 

At necropsy, the following changes in absolute and relative organ weight changes were noted:

absolute and relative weight of the hearts in females from the 10 mg/kg-day group were

significantly decreased; absolute brain weight of females from this same group were also

significantly decreased and relative group mean weight of the pituitary in males from the 3

mg/kg-day group was significantly increased. The biological significance of these weight

changes is difficult to assess, as they were not accompanied by morphologic changes. 

In animals that died before the end of the study, one male and two females from the 30 mg/kgday group and all animals from the 100 mg/kg-day group had marked diffuse lipid depletion in

the adrenal glands. All males and females from the 30 and 100 mg/kg-day groups also had slight

to moderate hypocellularity of the bone marrow and moderate atrophy of lymphoid follicles in

the spleen. One female from the 30 mg/kg-day group and all animals in the 100 mg/kg-day

group had moderate atrophy of the lymphoid follicles in the lymph nodes. 

Only one male in the 30 mg/kg-day group survived until terminal sacrifice, and this male also

had slight to moderate hypocellularity of the bone marrow and moderate atrophy of lymphoid

follicles in the spleen. No treatment related lesions were seen in the organs of animals from the

3 and 10 mg/kg-day groups.

The levels of PFOA in the serum and liver of these animals are presented below in Table 16. 

Individual values are presented so there are double entries for most dose levels.
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Table 16

Levels of PFOA in the Serum and Liver of Surviving Rhesus Monkeys

Dose

Serum (:g/ml) Liver (:g/ml) Liver total (:g)

Male Female Male Female Male Female

0 ND 1 0.05 0.07 3 5

3 53 65 3 7 250 350

3 48 50 ND ND ND ND

10 45 79 9 ND 600 ND

10 71 71 ND 10 ND 750

30 145 Dead 61 Dead 4000 Dead

ND - Not Determined

Under the conditions of this study the LOAEL is 3 mg/kg-day and no NOAEL was established

because effects were observed at the lowest dose.

Thomford (2001a,b) conducted a range-finding and a 6-month toxicity study in male

cynomolgus monkeys. In the range-finding study, Thomford (2001a) administered male

cynomolgus monkeys an oral capsule containing 0, 2, or 20 mg/kg-day APFO for 4 weeks. 

There were 3 monkeys in the 2 and 20 mg/kg-day groups and one monkey in the control group. 

The monkeys weighed 2.1 to 3.6 kg at the start of treatment. Animals were observed twice daily

for mortality and moribundity and were examined at least once daily for signs of poor health or

abnormal behavior. Body weights were recorded weekly and food consumption was assessed

qualitatively. The monkeys were fasted overnight and blood samples were collected one week

prior to the start of the study and on day 30 for clinical hematology and clinical chemistry

analyses, and hormone and PFOA level. Blood for clinical chemistry was also collected from

each animal on day 2 (approximately 24 hours after the first dose). Samples were analyzed for

estradiol, estrone, estriol, thyroid stimulating hormone, total and free triiodothyronine, and total

and free thyroxin. 

At scheduled necropsy, samples of the right lateral lobe of the liver were collected from each

animal and analyzed for palmitoyl CoA oxidase activity. Representative samples of liver, right

and left testes, and pancreas were collected from each animal for cell proliferation evaluation

using proliferation cell nuclear antigen. Bile was collected from each animal for bile acid

determination. A sample of liver was collected from each animal for PFOA concentration

analysis. The adrenals, liver, pancreas, spleen, and testes from each animal were examined

microscopically, and the remaining tissues were preserved for possible future examination.

All animals survived to scheduled sacrifice. There were no clinical signs of toxicity in the

treated groups and there was no effect on body weight. Low or no food consumption was

observed for one animal given 20 mg/kg-day. There were no effects on estradiol, estriol, thyroid

stimulating hormone, total and free triiodothyronine, and total and free thyroxin. Estrone levels

were notably lower for males given 2 and 20 mg/kg-day APFO. There was no evidence of
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50

peroxisome proliferation or cell proliferation in the liver, testes or pancreas of treated monkeys. 

No adverse effects were noted in either gross or clinical pathology studies. 

In the 26-week study, male cynomolgus monkeys were administered APFO by oral capsule at

doses of 0, 3, 10 or 30 mg/kg-day for 26 weeks (Thomford 2001b; Butenhoff et al., 2002). At

study initiation the monkeys weighed 3.2 to 4.5 kg. There were 4 monkeys in the 3 mg/kg-day

group and 6 monkeys in each of the other groups. Dosing of animals in the 30 mg/kg-day dose

group was stopped from days 11–21 because of toxicity. When dosing was resumed on day 22,

animals received 20 mg/kg-day and this group was designated the 30/20 mg/kg-day group. At

the end of the 26-week treatment period, 2 animals in the control and 10 mg/kg-day groups were

observed for a 13-week recovery period. 

Animals were observed twice daily for mortality and moribundity and were examined at least

once daily for signs of poor health or abnormal behavior. Ophthalmic examinations were done

before initiation of treatment and during weeks 26 and 40. Body weights were recorded weekly

and food consumption was assessed qualitatively. Blood and urine samples were collected for

clinical hematology, clinical chemistry, and urinalysis before the start of treatment and on days

11, 31, 63, 91, 182, 217, 245 and 275. Blood samples were also taken for hormone

determinations; samples were analyzed for estradiol, estrone, estriol, thyroid stimulating

hormone, total and free triiodothyronine, total and free thyroxin, testosterone, and

cholecystokinin (CCK). Blood, urine and feces were collected during week 2 and every 2 weeks

thereafter during treatment and recovery for PFOA concentration analyses. 

At scheduled necropsy, liver samples were taken for determination of PFOA levels. The right

lateral lobe of the liver was collected from each animal for palmitoyl CoA oxidase activity

analyses, and representative samples of liver, right and left testes, and pancreas were collected

from each animal for cell proliferation evaluation using proliferation cell nuclear antigen. All

available bile was collected for bile acid determination. Weights of the adrenal glands, brain,

epididymis, kidney, liver, pancreas, testis, and thyroid with parathyroid were recorded. The

following tissues were collected for histopathology: adrenal (2), aorta, brain, cecum, colon,

duodenum, epididymis (2), esophagus, eyes [preserved in Davidson’s fixative (2)], femur with

bone marrow (articular surface of the distal end), gallbladder, heart, ileum, jejunum, kidneys (2),

lesions, liver, lung, mesenteric lymph node, mammary gland, pancreas, pituitary, prostate,

rectum, salivary gland [mandibular (2)], sciatic nerve, seminal vesicle (2), skeletal muscle

(thigh), skin, spinal cord (cervical, thoracic, and lumbar), spleen, sternum with bone marrow,

stomach, testis [(2) preserved in Bouin’s solution], thymus, thyroid (2) with parathyroid, trachea

and urinary bladder. 

Two animals, one male from the 30/20 mg/kg-day dose group and one male from the 3 mg/kgday dose group, were sacrificed in moribund condition during the study. The male in the 30/20

mg/kg-day dose group was sacrificed on day 29. This animal exhibited signs of hypoactivity,

weight loss, few or no feces, low or no food consumption and the entire body was cold to the

touch before death. Necropsy revealed esophageal and gastric lesions that were indicative of an

injury that occurred during dosing and liver lesions that were presumed to be treatment related. 

The animal from the 3 mg/kg-day dose group was sacrificed on day 137. This animal showed

clinical signs of limited use and paralysis of the hind limbs, ataxia and hypoactive behavior, few

feces and no food consumption. The cause of death was not determined, but APFO treatment

could not be ruled out. 
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51

Males given 30 mg/kg-day from days 1-11 had clinical signs of few feces and low food

consumption and they lost weight during week 1 of treatment. Based on these signs, treatment

was stopped on day 11 and was not resumed until day 22. When treatment was resumed, the

dose was lowered to 20 mg/kg-day; this group was then designated the 30/20 mg/kg-day group. 

Of the remaining animals in this group, only 2 tolerated this dose level for the remaining 23

weeks of treatment. Treatment of three males given 30/20 mg/kg-day was halted on days 43

(week 7), 66 (week 10), and 81 (week 12) respectively. Clinical signs in these animals included

thin appearance, few or no feces, low or no food consumption, and weight loss. The animals

appeared to recover from compound-related effects within 3 weeks after cessation of treatment.

Mean body weight changes were notably lower during weeks 1 and 2 for males receiving 30

mg/kg-day. During week 2, this change was statistically significant. Treatment was stopped on

day 11 and when it was resumed at 20 mg/kg-day on day 21, mean body weight changes were

significantly lower than controls during weeks 7, 9 and 24. Overall mean body weight changes

through week 27 were notably lower for the males in the 30/20 mg/kg-day group. There was an

increased incidence of low or no food consumption for animals in the 30/20 mg/kg-day group

that was considered to be treatment related.

There were no consistent or clearly dose-related effects on estron, estradiol, estriol, testosterone,

or CCK levels in treated groups over time. Similarly, thyroid stimulating hormone and free and

total thyroxine levels remained relatively constant and were not significantly altered throughout

the study. Mean individual values for free and total triiodothyronine levels were statistically

significantly decreased during the dosing period at the 30/20 mg/kg/day dose group compared to

controls, although the biological significance of this decrease is unclear.

At terminal sacrifice at 26 weeks, there were statistically significant increases in mean absolute

liver weights and mean liver-to-body weight percentages in all dose groups. In addition, there

was a positve dose-response trend towards an increased relative liver-to-body weight, with

statistical significance at the high dose group of 30/20 mg/kg/day. The increased liver weights

were considered to be treatment-related. The increased liver weight was thought to be due, in

part, to hepatocellular hypertrophy (as demonstrated by decreased hepatic DNA content) which

in turn may be due to mitochodrial proliferation (as demonstrated by increased succinate

dehydrogenase activity). 

Since administration of APFO to rats results in liver, Leydig cell and pancreatic acinar cell

tumors, Butenhoff et al (2002) specifically looked for markers of tumor formation in the

monkeys. In the liver, there was only a two-fold increase in hepatic palmitoyl CoA oxidase

activity in the 30/20 mg/kg-day group, which is consistent with reports for other species that are

not particularly responsive to PPAR"-agonists. Replicative DNA synthesis in the liver, an

indication of cell proliferation, was not altered in the treated animals. Similarly, it has been

proposed that changes associated with the pancreatic acinar cell tumors in rats include increased

serum CCK concentrations and indications of cholestasis, including alkaline phosphatase,

bilirubin, and bile acids; none of these changes were noted in the cynomolgus monkeys. Finally,

in the rat, it has been proposed that the Leydig cell tumors are due to a sustained increase in

estradiol resulting from aromatase induction. In the treated cynomolgus monkeys there were no

significant changes in estradiol, estriol, or testosterone. In addition, there was no change in

replicative DNA synthesis in the pancreas or testes.

At the recovery sacrifice, there were no treatment-related effects on terminal body weights or on

absolute or relative organ weights indicating that the liver weight changes seen at terminal
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52

sacrifice were reversible over time. There were no treatment-related macroscopic or

microscopic changes at the recovery sacrifice.

Serum and liver concentrations of PFOA did not increase in a dose-related manner. This may

have been due to saturation or attaining steady state levels in the first several weeks of the study. 

In addition, there was a great deal of variability in the PFOA levels. This may have been due to

the method of dosing (by capsule), the timing of dosing relative to blood sample collection and

gall bladder emptying, or the analytical method (+ 30% for interassay, intra-assay and system). 

Since steady state appeared to have been reached by 4-6 weeks of dosing, the study authors

calculated the mean serum levels for the period following 6 weeks; steady state serum levels

were 77 + 39, 86 + 33 and 158 + 100 :g/ml for the 3, 10 and 30/20 mg/kg-day groups,

respectively. In the control animals, 2/3 of the serum samples contained PFOA and averaged a

level of 0.203 + 0.154 :g/ml. There was not a statistically significant difference in the mean

serum levels of the 3 and 10 mg/kg-day groups, but the mean serum level in the 30/20 mg/kgday group was significantly higher than the 3 and 10 mg/kg-day groups. At terminal sacrifice,

the levels of PFOA in the liver were similar in the 3 and 10 mg/kg-day groups, and ranged from

6.29 - 21.9 :g/g. The two monkeys in the 30/20 mg/kg-day group had liver concentrations of 16

and 83.3 :g/g. After the recovery period, the serum and liver PFOA levels in the 10 mg/kg-day

group had returned to baseline. 

Under the conditions of the study, the LOAEL was 3 mg/kg-day (increased liver weight and

possibly mortality) and a NOAEL was not established.

3.5.2 Subchronic Studies in Rodents

Christopher and Marisa (1977) administered ChR-CD mice (5/sex/group) dietary concentrations

of 0, 30, 100, 300, 1000, 3000, 10,000, or 30,000 ppm of APFO for 28 days. The animals were

observed daily and body weights and food consumption were recorded weekly. At necropsy, the

organs were weighed, examined for gross pathology and preserved for histopathology. All

animals in the 1000 ppm and higher groups died before the end of day 9. All animals in the 300

ppm group died within 26 days except one male. One animal in each of the 30 and 100 ppm

groups died prematurely. Clinical signs were observed in mice exposed to 100 ppm and higher

doses of APFO. At 100 ppm some animals exhibited cyanosis on days 10 and 11 of testing, but

appeared normal throughout the rest of the study. Animals fed 300 ppm exhibited roughed fur

and muscular weakness as well as signs of cyanosis after 9 days of treatment. Animals fed 1000

ppm exhibited similar effects after 6 days and those receiving 3000 ppm or greater doses

exhibited effects after 4 days. There was a dose-related reduction in mean body weight in all

treated groups. Relative and absolute liver weights were increased in mice fed 30 ppm or more

APFO. Treatment-related changes were observed in the livers among all APFO treated animals

including enlargement and/or discoloration of 1 or more liver lobes. Histopathologic

examination of all surviving treated mice revealed diffuse cytoplasmic enlargement of

hepatocytes throughout the liver (panlobular hypertrophy) accompanied by focal to multifocal

cytoplasmic lipid vacuoles of variable size which were random in distribution. 

Metrick and Marias (1977) administered ChR-CD rats (5/sex/group) dietary concentrations of 0,

30, 100, 300, 1000, 3000, 10,000, or 30,000 ppm APFO for 28 days. The animals were observed

daily and body weights and food consumption were recorded weekly. At necropsy, the organs

were weighed, examined for gross pathology and preserved for histopathology. All animals in

the 10,000 and 30,000 ppm groups died before the end of the first week. Gross pathologic

examination revealed white foci in the cortex and medulla in the kidneys of a 10,000 ppm
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53

female. Pelvic dilation was evident in the kidneys of a control male and a 30,000 ppm female. 

There were no premature deaths or unusual behavioral reactions in the other groups. Food

consumption was reduced in the 1000 and 3000 ppm groups. Body weight gain was reduced as

dose increased. The reduction in body weight gain was statistically significant for males at 1000

ppm and males and females at 3000 ppm. Absolute liver weights were increased in males fed 30

ppm or more and females fed 300 ppm or more. Treatment-related morphologic changes were

observed in the livers of all test animals. Focal to multifocal cytoplasmic enlargement of

hepatocytes in the centrilobular to midzonal areas was noted in animals fed 30 to 300 ppm, and 

multifocal to diffuse enlargement of hepatocytes throughout the liver lobules (panlobular) was

noted in animals fed 1000 ppm or higher. The hypertrophy of hepatocytes was accompanied by

acidophilic degeneration and/or necrosis of scattered liver cells with no lobular distribution. The

severity and degree of tissue involvement were more pronounced in males than in females. 

Goldenthal (1978a) administered ChR-CD rats (5/sex/group) dietary levels of 0, 10, 30, 100,

300, and 1000 ppm APFO for 90 days. These dose levels are equivalent to 0.056, 1.72, 5.64,

17.9, and 63.5 mg/kg-day in males, and 0.74, 2.3, 7.7, 22.36 and 76.47 mg/kg-day in females. 

Animals were observed twice daily and body weight and food consumption were recorded

weekly. Blood and urine samples were collected during the pretest period and at 1 and 3 months

of the study for hematology and clinical chemistry and urinalysis. At necropsy, the organs from

the control, 100, 300, and 1000 ppm groups were weighed and examined for histopathologic

lesions; livers from the 10 and 30 ppm groups were also examined microscopically.

There were no treatment-related changes in behavior or appearance. One female in the 100 and

one female in the 300 ppm group died during collection of blood. These deaths were not

considered to be treatment related. All other animals survived until scheduled sacrifice. There

was a decrease in body weight gain for male rats at the 300 and 1000 ppm dose level. At 13

weeks, mean body weight of males in the 1000 ppm group was significantly less than that of

controls. There were no treatment related effects on the hematologic, biochemical or urine

parameters.

Relative kidney weights were significantly increased in males in the 100, 300, and 1000 ppm

groups. However, absolute kidney weights were comparable among groups, and there were no

histopathological lesions. Absolute liver weights were significantly increased in males in the 30,

300 and 1000 ppm groups and in females in the 1000 ppm group. Relative liver weights were

significantly increased in males in the 300 and 1000 ppm groups and in females in the 1000 ppm

group. Discoloration on the surface of the liver was observed in male rats in the 1000 ppm

group. Hepatocellular hypertrophy (focal to multifocal in the centrilobular to midzonal regions)

was observed in 4/5, 5/5, and 5/5 males in the 100, 300, and 1000 ppm groups, respectively. 

Hepatocyte necrosis was observed in 2/5, 2/5, 1/5, and 2/5 males in the 30, 100, 300, and 1000

ppm groups, respectively. Under the conditions of this study, the LOAEL for males is 30 ppm

(1.72 mg/kg-day) based on liver effects and the NOAEL is 10 ppm (0.56 mg/kg-day); the

LOAEL for females is 1000 ppm (76.5 mg/kg-day) and the NOAEL is 300 ppm (22.4 mg/kgday).

Palazzolo (1993) administered male ChR-CD rats (45-55 per group) dietary concentrations of 1,

10, 30, or 100 ppm APFO for 13 weeks. These doses are equivalent to 0.06, 0.64, 1.94, and 6.50

mg/kg-day. Two control groups (a nonpair-fed control group and a control group pair-fed to the

100 ppm dose group) were fed basal diet during that period. Following the 13-week exposure

period, 10 animals per group were fed basal diet for an 8-week recovery period. The animals

were observed twice daily for clinical signs of toxicity, and body weights and food consumption
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54

were recorded weekly. Food consumption was recorded daily for the pair-fed animals. A total

of 15 animals per group were sacrificed following 4, 7, or 13 weeks of treatment; 10 animals per

group were sacrificed after 13 weeks of treatment and following the 8 week recovery period. 

Serum samples collected from 10 animals per group at each scheduled sacrifice during treatment

and from 5 animals per group during recovery were analyzed for estradiol, total testosterone,

luteinizing hormone, and PFOA. The level of palmitoyl CoA oxidase was analyzed from a

section of liver that was obtained from 5 animals per group at each scheduled sacrifice. Weights

of the brain, liver, lungs, testis, seminal vesicle, prostate, coagulating gland, and urethra were

recorded, and these tissues were examined histologically. In addition, the brain, liver, lungs,

testis, seminal vesicle, and prostate were preserved in glutaraldehyde for electron microscopic

examination.

In the analysis of the data, animals in groups exposed to 1, 10, 30, and 100 ppm APFO were

compared to the control animals in the nonpair-fed group, while the data from the pair-fed

control animals were compared to animals exposed to 100 ppm APFO. No treatment-related

clinical signs were noted. At 100 ppm, significant reductions in body weights were seen

compared to the pair-fed control group during week 1 and the nonpair-fed control group during

weeks 1-13. Body weight data in the other dosed-groups were comparable to controls. At 100

ppm, mean body weight gains were significantly higher than the pair-fed control group during

week 1 and significantly lower than the nonpair-fed control group during weeks 1-13. At 10 and

30 ppm, mean body weight gains were significantly lower than the nonpair-fed control group at

week 2. These differences in body weight and body weight gains were not observed during the

recovery period. Animals fed 100 ppm consumed significantly less food during weeks 1 and 2,

when compared to the nonpair-fed control group. Overall, there was no significant difference in

food consumption. There were no significant differences among the groups for any of the

hormones evaluated in the serum although there appeared to be some indication of elevated

estradiol for the 100 ppm group at week 5. 

Significant increases in absolute and relative liver weights and hepatocellular hypertrophy were

observed at weeks 4, 7, or 13 in the 10, 30 and 100 ppm groups. There was no evidence of any

degenerative changes or abnormalities associated with the hypertrophy. Hepatic palmitoyl CoA

oxidase activity was significantly increased at weeks 4, 7, and 13 in the 30 and 100 ppm groups. 

At 10 ppm, hepatic palmitoyl CoA oxidase activity was significantly increased at week 4 only. 

During recovery, however, none of the liver effects were observed, indicating that these

treatment-related liver effects were reversible.

Serum levels from the male rats in this study are shown below in Table 17.
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55

Table 17

Summary of Serum Values (:g/ml) in Male Rats Treated with APFO for 90 Days

Week

Dose Group (mg/kg-day)

0.06 0.64 1.94 6.50

5 6.5 ± 1.05a

 (8)

55.4 ± 8.08

 (9)

103.5 ± 14.38

 (8)

159.3 ± 30.16

 (10)

8 7.5 ± 1.27

 (9)

45.5 ± 16.20

 (10)

87.3 ± 27.48

 (10)

149.1 ± 34.98

 (10)

14 7.1 ± 1.15

 (10)

41.2 ± 12.98

 (10)

70.3 ± 16.18

 (10)

137.6 ± 33.83

 (10)

22 1.2 

(1)

1.09 ± 1.303

 (3)

1.64 ± 0.918

 (3)

2.45 ± 0.849

 (2)

a - Mean + SD, (N)

Under the conditions of this study, the LOAEL is 10 ppm (0.64 mg/kg-day) based on increases

in absolute and relative liver weights with hepatocellular hypertrophy. The NOAEL is 1.0 ppm

(0.06 mg/kg-day).

3.5.3 Chronic Toxicity and Carcinogenicity Studies in Rats

The chronic toxicity and carcinogenic potential of PFOA has been investigated in two dietary

studies in rats. In the first study (Sibinski, 1987), groups of 50 male and 50 female SpragueDawley (Crl:CD BR) rats were fed diets containing 0, 30 or 300 ppm APFO for two years.

Groups of 15 additional rats per sex were fed 0 or 300 ppm APFO and evaluated at the one year

interim sacrifice. In males. the mean test article consumption was 1.3 and 14.2 mg/kg-day for

the 30 and 300 ppm groups, respectively; in females, the mean test article consumption was 1.6

and 16.1 mg/kg-day for the 30 and 300 ppm groups, respectively. All animals were observed

daily throughout the two year dosing period. Body weights and feed consumption were recorded

once per week for the first six months, and then once every two weeks for the remainder of the

study. Clinical pathological examinations including hematology, serum chemistry and urinalysis

were conducted on samples obtained from 15 rats per sex from each group at 3, 6, 12, 18 and 24

months. Macroscopic postmortem examinations were performed on all animals that died during

the study and those which were terminated at the one year interim and two year necropsies. The

weights of the kidneys, liver, testes, brain, heart, spleen, adrenal glands and uterus were recorded

for 15 randomly selected rats/sex at the interim termination from both the control and high-dose

groups, and from the control and both treated groups at the two year necropsy. Microscopic

evaluation was performed on all tissues from all of the control and high-dose rats.

There was a dose-related decrease in body weight gains in the males rats and to a lesser extent,

in the female rats as compared to the controls; the decreases were statistically significant in the

high-dose groups of both sexes (up to 21% in male rats and 11% in female rats). The body

weight changes are treatment related since feed consumption was actually increased (rather than

decreased). There were no differences in mortality between the treated and untreated groups; the
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56

survival rates at the end of 104 weeks for the male control, low-, and high-dose groups were

70%, 72% and 88%, respectively; in females, the survival rates were 50%, 48% and 58% for the

control, 30 and 300 ppm groups, respectively. The only clinical sign observed was a doserelated increase in ataxia in the female rats; which was most commonly associated with

moribund animals; the incidences in the control, low- and high-dose groups were 4%, 18% and

30%. Significant decreases in red blood cell counts, hemoglobin concentrations and hematocrit

values were observed in the high-dose male and female rats as compared to control values. 

Clinical chemistry changes included slight (less than 2-fold), but significant increases (P< 0.05)

in alanine aminotransferase (ALT), aspartate aminotransferase (AST) and alkaline phosphatase

(AP) in both treated male groups from 3-18 months, but only in the high-dose males at 24

months. Slight (up to about 10%) increases in absolute or relative liver and kidney weights were

noted in both high-dose male and female rats at the 1 year interim sacrifice and at the terminal

necropsy; however, only the relative liver weight (vs. body weight or brain weight) increases in

the high-dose males were statistically significant (P<0.05). 

Histologic evaluations showed lesions in the liver, testis and ovary. In the liver, the increased

incidence of lesions reached statistical significance only in the high-dose male group. At the 1

year interim sacrifice, diffuse hepatomegalocytosis (12/15 animals), portal mononuclear cell

infiltration (13/15 animals) and hepatocellular necrosis (6/15 animals) were seen in the highdose males while incidences in the control group were, 0/15, 7/15 and 0/15, respectively. 

Hepatocellular vacuolation was seen in 11/15 high-dose females as compared to an incidence of

5/15 in the control group. At the 2-year sacrifice, megalocytosis was found at an incidence of

0%, 12% and 80% in the males, and 0%, 2% and 16% in the females from the control, low-, and

high-dose groups, respectively. Hepatic cystoid degeneration, a condition characterized by areas

of multilocular microcysts in the liver parenchyma, was observed in 14% and 56% of the lowand high-dose males, as compared to a control incidence of 8%. The incidence of hyperplastic

nodules, a localized proliferation of hepatic parenchymal cells, was slightly increased in the

high-dosed males with an incidence of 6% as compared to 0% in the control males.

 

At the one-year sacrifice, testicular masses were found in 6/50 high-dose and 1/50 low-dose rats,

but not in any of the controls. Furthermore, marked aspermatogenesis was found in 2/15 highdose males but none in the control males. At the 2-year sacrifice, vascular mineralization of the

testes occurred in 18% of the high-dosed male and 6% of the low-dosed males, but was not seen

in the controls. These testicular effects reached statistical significance in the high-dose group. 

A statistically significant, dose-related increase in the incidence of ovarian tubular hyperplasia

was found in female rats at the 2-year sacrifice. The incidence of this lesion in the control, low-,

and high-dose groups was 0%, 14%, and 32%, respectively. The biological significance of this

effect at the time of the initial evaluation was unknown, as there was no evidence of progression

to tumors. Recently, however, slides of the ovaries from that study were re-evaluated, with

particular emphasis placed on the proliferative lesions of the ovary (Mann and Frame, 2004). 

Using more recently published nomenclature, the ovarian lesions were diagnosed and graded as

gonadal stromal hyperplasia and/or adenomas, which corresponded to the diagnoses of tubular

hyperplasia or tubular adenoma by the original study pathologist. The data are summarized in

Table 18. No statistically significant increases in hyperplasia (total number), adenomas, or

hyperplasia/adenoma combined were seen in treated groups compared to controls. There was

some evidence of an increase in size of stromal lesions observed at the 300 ppm group; however,

adenomas occurred in greater incidences in the control group than in either of the treated groups. 

Results of this follow-up evaluation indicated that rats sacrificed at the one-year interim
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57

sacrifice, as well as rats that died prior to the interim sacrifice were not considered at risk for

tumor development. 

Table 18

Incidence of Ovarian Stromal Hyperplasia and Adenoma in Rats

Group Control 30 ppm 300 ppm

No. examined 45 47 46

Hyperplasia (Total) 8 16 15

 Grade 1 6 7 5

 Grade 2 2 3 1

 Grade 3 0 5 6

 Grade 4 0 1 3

Adenoma 4 0 2

Adenoma and/or

Hyperplasia 12 16 17

From Mann and Frame (2004)

Based on these toxic effects, the high dose selected in this study appears to have reached the

Maximum Tolerated Dose. Based on a decrease in body weight gain, increase in liver and

kidney weights and toxicity in the hematological and hepatic systems, the LOAEL for male rats

is 300 ppm and the NOAEL is 30 ppm. The LOAEL for female rats is 300 ppm based on a

decrease in body weight gain and hematologic effects and the NOAEL is 30 ppm. 

At the termination of the study, there was a significant increase (P<0.05) in the incidence of

testicular (Leydig) cell adenomas in the high-dose male rats. The incidence of the Leydig cell

tumors (LCT) in the control, low- and high-dose males was 0/50 (0%), 2/50 (4%) and 7/50

(14%), respectively. The increase was also statistically significant when compared to the

historical control incidence of 0.82% observed in 1,340 Sprague-Dawley control male rats used

in 17 carcinogenicity studies (Chandra et al., 1992). The spontaneous incidence of LCT in 2-

year old Sprague-Dawley rats in other studies was reported to be approximately 5% (cited in

Clegg et al., 1997). 

There was also a significant increase (P<0.05) in the incidence of mammary fibroadenomas in

both groups of female rats. The incidence of the mammary fibroadenoma was 10/47 (21%),

19/47 (40%) and 21/49 (43%) in the control, 30, and 300 ppm groups, respectively. The increase

was also statistically significant when compared to the historical control incidence of 19.0%

observed in 1,329 Sprague-Dawley control female rats used in 17 carcinogenicity studies

(Chandra et al., 1992). The investigators did not consider the mammary fibroadenomas to be

treatment related on the basis of the historical control incidence (24%) from a study of 181

female rats terminally sacrificed at 18 months, which is considered an inappropriate historical

reference. When the mammary fibroadenoma incidences were compared to the historical control
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58

incidence (37%) in 947 female rats in the Haskell Laboratory, however, there does not appear to

be any compound related effect (Sykes,1987).

The induction of Leydig cell tumors was confirmed in a follow-up 2-year mechanistic study of

PFOA toxicity in male Sprague-Dawley rats at a dietary level of 300 ppm (Cook et al., 1994;

Biegel et al., 2001). There were 156 animals in the treatment group and 80 animals in the

control group. Cage-side examinations were conducted at least once daily throughout the study. 

Rats were weighed once a week during the first 3 months and once every other week for the

remainder of the study. Rats were euthanized at interim time points of 1, 3, 6, 9, 12, 15, 18, and

21 months. At each time point, the liver and testis from 6 rats/group were weighed and

evaluated for cell proliferation. Another 6 rats/group were selected for peroxisome proliferation,

and 10 rats/group for serum hormone (estradiol, testosterone, LH, FSH, and prolactin) analysis. 

All rats surviving the 24-month test period were necropsied for microscopic examination of

various organs: e.g., kidneys, liver, testes, brain, heart, spleen. 

In the treated group, relative liver weights and hepatic $-oxidation activity were statistically

significantly increased at all of the sampling time points when compared to the controls. 

Absolute testis weights were increased only at 24 months. No hepatic or Leydig cell

proliferation was observed at any of sampling times. There were no significant differences in

serum testosterone, FSH, LH, or prolactin in the PFOA-treated rats when compared to the

controls. There were, however, significant increases in serum estradiol concentrations in the

treated rats at 1, 3, 6, 9, and 12 months.

There was a significant increase in the incidence of LCT in the treated rats (8/76, 11%) as

compared to the controls (0/80, 0%). In addition, the treated group had a significant increase in

the incidences of liver adenomas and pancreatic acinar cell tumors (PACT). The incidences of

liver adenomas in the control and treated groups were 2/80 (3%) and 10/76 (13%), respectively,

whereas those for the pancreatic acinar cell adenomas were 0/80 (0%) and 7/76 (9%). There was

one pancreatic acinar cell carcinoma in 76 of the treated rats and none in 80 controls. The

incidence of combined pancreatic acinar cell adenoma/carcinoma in the treated rats (8/76, 11%)

was significantly increased as compared with the controls (0/80, 0%).

In the first carcinogenicity study (Sibinski, 1987), there was no reported increase in the

incidence of PACT, and the incidence of pancreatic acinar hyperplasia in the male rats was 0/33,

2/34, and 1/43 in the control, 30 and 300 ppm groups, respectively. To resolve this discrepancy,

the histological slides from both studies were reviewed by independent pathologists. This

review of the microscopic lesions of the pancreas in the two studies indicate that PFOA

produced increased incidences of proliferative acinar cell lesions of the pancreas in the rats of

both studies at the dietary concentration of 300 ppm. The differences observed were quantitative

rather than qualitiative; more and larger focal proliferative acinar cell lesions and greater

tendency for progression of lesions to adenoma of the pancreas were observed in the second

study compared to the first study. The difference between pancreatic acinar hyperplasia

(reported in Sibinski, 1987) and adenoma (reported in Cook et al., 1994; Biegel et al. 2001) in

the rat is a reflection of arbitrary diagnostic criteria and nomenclature by different pathologists. 

The basis for the quantitative difference in the lesions observed in not known but was believed to

be due most likely to difference in the diets used in the two laboratories (Frame and McConnell,

2003). 

In summary, the two carcinogenicity studies of PFOA have shown that PFOA induced liver

adenomas, Leydig cell adenomas, and pancreatic acinar cell tumors in male Sprague-Dawley 
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59

rats. The evidence for mammary fibroadenomas in the female rats is equivocal since the

incidences were comparable to some historical background incidences. PFOA has also been

shown to promote liver carcinogenesis in rodents (Abdellatif et al.,1991; Nilsson et al., 1991).

3.6 Immunotoxicity Studies in Mice

Four immunotoxicity studies of PFOA have been conducted in mice. In the first study, Yang et

al. (2000) administered 0.02 % PFOA to male C57Bl/6 mice in the diet for 2, 5, 7, or 10 days. 

At the end of the feeding period, mice were sacrificed and the liver, spleen, and thymus were

weighed. The effect of PFOA administration on the cellularity, cell surface phenotype, and cell

cycle of thymocytes and splenocytes was determined. In addition, effects of exposure of

thymocytes and splenocytes to PFOA in vitro were examined. Administration of 0.02% PFOA

for 2, 5, 7, or 10 days resulted in a significant increase, relative to controls, in liver weight, even

at the earliest time point. Also, a decrease in body weight was observed. Following five days of

administration, significant decreases in thymus and spleen weight were noted. After

administration of 0.02% PFOA for 7 days, significant decreases (85% and 80%, respectively) in

the total number of thymocytes and splenocytes were observed. In addition, the number of

thymocytes expressing both CD4 and CD8 decreased by 95%; the number expressing both CD4

and CD8 decreased by 57%; and the number expressing either CD4 or CD8 decreased by 64%

and 72%, respectively. For the splenocytes, both T cells (CD3) and B cells (CD19) decreased by

75% and 86%, respectively. Also, significant decreases in both CD4 helper and CD8 cytotoxic

splenic T cells were observed. Upon administration of 0.02% PFOA to mice for 7 days,

thymocyte proliferation was also inhibited, as detected by cell cycle flow cytometry analyses. In

vitro studies showed that there was spontaneous apoptosis occurring in splenocytes and

thymocytes after 8 or 24 hours of culturing in the presence of varying concentrations (50, 100, or

200 M) of PFOA. However, PFOA did not significantly alter the cell cycle under these

conditions. 

In order to examine the dose dependency of the effects, Yang et al. (2001) administered C57Bl/6

mice diets consisting of 0.001%-0.05% PFOA (w/w) for 10 days. For examining the timecourse, a diet containing 0.02% PFOA was given for 2, 5, 7 or 10 days. Effects of withdrawal of

PFOA were also studied. The results showed that, at higher doses, a significant decrease,

relative to controls, in body weight was observed, although no other apparent signs of toxicity

such as sores, lethargy, and poor grooming were noticed. However, a significant decrease in

total water intake was observed. Mice receiving dietary PFOA for 10 days experienced

significant increases in liver weight and peroxisome proliferation, as measured by induction of

acyl-CoA oxidase with lauroyl-CoA or palmitoyl-CoA as substrate. These increases started at

the lowest dose and reached their maximal values at a dose of 0.003-0.01%. In contrast, the

weight decreases of the spleen and thymus began at a higher dose (0.01%) with no maximum

reached with the doses given. The time course studies showed that increased liver weights and

peroxisome proliferation were evident at the earliest time point examined. In contrast,

significant thymus and spleen weight decreases required PFOA administration for a period of at

least 5 days, following which the spleen weight remained constant while the thymus weight

continued to decrease. However, upon prolonged treatment for one month, no further decreases

in thymus and spleen weights were observed. In another set of experiments, animals received

0.02% PFOA for 7 days, and then they received normal chow for a period of 10 days. These

recovery experiments showed that the animals rapidly recovered the body weight the second day

after withdrawal of PFOA. However, the liver weight did not return to normal even after 10

days of recovery. Thymus recovery started on day 2 and was completed by day 10. The spleen

weights returned to normal by day 2 post-withdrawal. In addition, the changes in thymus and
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spleen weight upon PFOA treatment and withdrawal paralleled the changes in total thymocyte

and splenocyte counts. Furthermore, flow cytometry cell cycle experiments showed that the

decrease in thymocyte number caused by PFOA treatment is due mainly to inhibition of

thymocyte proliferation. In contrast, PFOA treatment caused no changes in the cell cycle of

splenocytes.

 

A third feeding study (Yang et al. 2002a) was designed to examine the possible involvement of

the peroxisome proliferator-activated receptor alpha (PPARα) in the immunomodulation exerted

by PFOA. This study made use of transgenic PPARα null mice, which are homozygous with

regards to a functional mutation in the PPARα gene. These mice do not exhibit peroxisome

proliferation or hepatomegaly and hepatocarcinogenesis even after exposure to peroxisome

proliferators. These mice were fed a diet consisting of 0.02% PFOA (w/w) for 7 days. At the

end of the feeding period, mice were sacrificed and the liver, spleen, and thymus were removed

and weighed. The effect of PFOA on peroxisome proliferation, cell cycle, and

lymphoproliferation was ascertained. 

The results showed that, in contrast to wild-type mice, feeding PPARα null mice PFOA resulted

in no significant decrease in body weight. As expected, peroxisome proliferation, as measured

by fatty acid oxidation, was totally lacking in PPARα null mice. Also in contrast to wild type

mice, feeding PPAR" null mice PFOA resulted in no significant decrease in the weight of the

spleen or the number of splenocytes. At the same time, the decrease in weight and cellularity of

the thymus was attenuated, but not totally eliminated in the PPAR" null mice. In addition, the

decreases in the size of the CD4+CD8+ population of thymus cells and the number of thymus

cells in the S and G2/M phases of the cell cycle, which reflects inhibition of proliferation,

observed in wild type mice administered PFOA were much less extensive in PPAR" null mice. 

Finally, in contrast to wild type mice, PFOA treatment caused no significant change in

splenocyte proliferation in response to mitogens in PPAR" null mice. 

A fourth feeding study (Yang et al. 2002b) was designed to examine the effects of PFOA on

specific humoral immune responses in mice. For this study, 0.02 % PFOA was administered to

male C57Bl/6 mice for 10 days. Then the animals were examined, via plaque forming cell

(PFC) and serum antibody assays, for their ability to generate an immune response to horse red

blood cells (HRBCs). Ex vivo and in vitro splenic lymphocyte proliferation assays were also

performed. The results showed that mice fed normal chow responded to challenge with HRBCs

with a strong humoral response, as measured by the PFC assay. In contrast, mice fed with PFOA

responded to HRBC immunization with no increase in HRBC-specific PFCs, relative to

unimmunized controls. However, in experiments where PFOA-treated mice received normal

chow following HRBC immunization, there was a significant recovery of the numbers of

specific PFCs stimulated. The suppression of the humoral immune response by PFOA was

confirmed by analysis of the serum anti-HRBC response. In ex vivo experiments, splenocytes

isolated from control mice responded to both ConA and LPS with lymphocyte proliferation, as

measured by thymidine incorporation. However, treating mice with PFOA (0.02% for 7 days)

attenuated the proliferation. In a set of in vitro experiments, PFOA (1- 200 M) added to the

culture medium of splenocytes cultured from untreated mice did not cause an alteration of

lymphocyte proliferation in response to LPS or ConA. 

3.7 Prenatal Developmental Toxicity Studies in Animals

Several prenatal developmental toxicity studies of APFO have been conducted. These include

two oral studies in rats, one oral study in rabbits, and one inhalation study in rats. 
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Gortner (1981) administered time-mated Sprague-Dawley rats (22 per group) doses of 0, 0.05,

1.5, 5, and 150 mg/kg-day APFO in distilled water by gavage on gestation days (GD) 6-15. 

Doses were adjusted according to body weight. Dams were monitored on GD 3-20 for clinical

signs of toxicity. Individual body weights were recorded on GD 3, 6, 9, 12, 15, and 20. Animals

were sacrificed on GD 20 by cervical dislocation and the ovaries, uteri, and contents were

examined for the number of corpora lutea, number of viable and non-viable fetuses, number of

resorption sites, and number of implantation sites. Fetuses were weighed and sexed and

subjected to external gross necropsy. Approximately one-third of the fetuses were fixed in

Bouin’s solution and examined for visceral abnormalities by free-hand sectioning. The

remaining fetuses were subjected to skeletal examination using alizarin red.

Signs of maternal toxicity consisted of statistically significant reductions in mean maternal body

weights on GD 9, 12, and 15 at the high-dose group of 150 mg/kg-day. Mean maternal body

weight on GD 20 continued to remain lower than controls, although the difference was not

statistically significant. Other signs of maternal toxicity that occurred only at the high-dose

group included ataxia and death in three rat dams. No other effects were reported. 

Administration of APFO during gestation did not appear to affect the ovaries or reproductive

tract of the dams. Under the conditions of the study, a NOAEL of 5 mg/kg-day and a LOAEL of

150 mg/kg-day for maternal toxicity were indicated.

A significantly higher incidence in fetuses with one missing sternebrae was observed at the highdose group of 150 mg/kg-day; however this skeletal variation also occurred in the controls and

the other three dose groups (at similar incidence but lower than the high-dose group) and

therefore was not considered to be treatment-related. No significant differences between treated

and control groups were noted for other developmental parameters that included the mean

number of males and females, total and dead fetuses, the mean number of resorption sites,

implantation sites, corpora lutea and mean fetus weights. Likewise, a fetal lens finding initially

described as a variety of abnormal morphological changes localized to the area of the embryonal

nucleus, was later determined to be an artifact of the free-hand sectioning technique and

therefore not considered to be treatment-related. Under the conditions of the study, a NOAEL

for developmental toxicity of 150 mg/kg-day (highest dose group) was indicated.

A second oral prenatal developmental toxicity study was conducted in rabbits (Gortner, 1982). 

Based on the results of a range-finding study, an upper dose level of 50 mg/kg-day was set for

the definitive study in which four groups of 18 pregnant New Zealand White rabbits were

administered 0, 1.5, 5, and 50 mg/kg-day APFO in distilled water by gavage on gestation days 6-

18. Pregnancy was established in each sexually mature female by i.v. injection of pituitary

lutenizing hormone in order to induce ovulation, followed by artificial insemination with 0.5 ml

of pooled semen collected from male rabbits; the day of insemination was designated as day 0 of

gestation. A constant dose volume of 1 ml/kg was administered. Individual body weights were

measured on GD 3, 6, 9, 12, 15, 18, and 29. The does were observed daily on GD 3-29 for

abnormal clinical signs. On GD 29, the does were euthanized and the ovaries, uterus and

contents examined for the number of corpora lutea, live and dead fetuses, resorptions and

implantation sites. Fetuses were examined for gross abnormalities and placed in a 370 C

incubator for a 24-hour survival check. Pups were subsequently euthanized and examined for

visceral and skeletal abnormalities. 

 

Signs of maternal toxicity consisted of statistically significant transient reductions in body

weight gain on GD 6-9 when compared to controls; body weight gains returned to control levels
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on GD 12-29. Administration of APFO during gestation did not appear to affect the ovaries or

reproductive tract contents of the does. No clinical or other treatment-related signs were

reported. Under the conditions of the study, a NOAEL of 50 mg/kg-day, the highest dose tested,

for maternal toxicity was indicated. 

No significant differences were noted between controls and treated groups for the number of

male and female fetuses, dead or live fetuses, or fetal weights. Likewise, there were no

significant differences reported for the number of resorption and implantation sites, corpora

lutea, the conception incidence, abortion rate, or the 24-hour mortality incidence of the fetuses. 

Gross necropsy and skeletal/visceral examinations were unremarkable. The only sign of

developmental toxicity consisted of a dose-related increase in a skeletal variation, extra ribs or

13th rib, with statistical significance at the high-dose group (38% at 50 mg/kg-day, 30% at 5

mg/kg-day, 20% at 1.5 mg/kg-day, and 16 % at 0 mg/kg-day). A statistically significant increase

in 13th ribs-spurred occurred in the mid-dose group of 5 mg/kg-day; however, the biological

significance of this effect is uncertain since in both the high- and low-dose groups, this effect

occurred at the same rate and was not statistically significantly different from controls. 

Therefore, under the conditions of the study, a LOAEL for developmental toxicity of 50 mg/kgday (highest dose group) was indicated.

 

Staples et al. (1984) also conducted a developmental toxicity study of APFO in rats. The study

design consisted of an inhalation and an oral portion, each with two trials or experiments. In the

first trial the dams were sacrificed on GD 21; while in the second trial, the dams were allowed to

litter and the pups were sacrificed on day 35-post partum. For the inhalation portion of the

study, the two trials consisted of 12 pregnant Sprague-Dawley rats per group exposed to 0, 0.1,

1, 10, and 25 mg/m3 APFO by whole-body dust inhalation for 6 hours/day, on GD 6-15. In the

oral portion of the study, 25 and 12 Sprague-Dawley rats for the first and second trials,

respectively, were administered 0 and 100 mg/kg-day APFO in corn oil by gavage on GD 6-15. 

For both routes of administration, females were mated on an as-needed basis and when the

number of mated females was bred, they were ranked within breeding days by body weight and

assigned to groups by rotation in order of rank. Finally, two additional groups (six dams per

group) were added to each trial that was pair-fed to the 10 and 25 mg/m3 groups.

For trial one, the dams were weighed on GD 1, 6, 9, 13, 16, and 21 and observed daily for

abnormal clinical signs. On GD 21, the dams were sacrificed by cervical dislocation and

examined for any gross abnormalities, liver weights were recorded and the reproductive status of

each animal was evaluated. The ovaries, uterus and contents were examined for the number of

corpora lutea, live and dead fetuses, resorptions and implantation sites. Pups (live and dead)

were counted, weighed and sexed and examined for external, visceral, and skeletal alterations.

The heads of all control and high-dosed group fetuses were examined for visceral alterations as

well as macro- and microscopic evaluation of the eyes. 

For trial two, in which the dams were allowed to litter, the procedure was the same as that for

trial one up to GD 21. Two days before the expected day of parturition, each dam was housed in

an individual cage. The date of parturition was noted and designated Day 1 PP. Dams were

weighed and examined for clinical signs on Days 1, 7, 14, and 22 PP. On Day 23 PP all dams

were sacrificed. Pups were counted, weighed, and examined for external alterations. Each pup

was subsequently weighed and inspected for adverse clinical signs on Days 4, 7, 14, and 22 PP. 

The eyes of the pups were also examined on Days 15 and 17 PP for the inhalation portion and on

Days 27 and 31 PP for the gavage portion of the study. Pups were sacrificed on Day 35 PP and

examined for visceral and skeletal alterations.
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In trial one of the inhalation study, treatment-related clinical signs of maternal toxicity occurred

at 10 and 25 mg/m3

 and consisted of wet abdomens, chromodacryorrhea, chromorhinorrhea, a

general unkempt appearance, and lethargy in four dams at the end of the exposure period (highconcentration group only). Three out of 12 dams died during treatment at 25 mg/m3

 (on GD 12,

13, and 17). Food consumption was significantly reduced at both 10 and 25 mg/m3

; however, no

significant differences were noted between treated and pair-fed groups. Significant reductions in

body weight were also observed at these concentrations, with statistical significance at the highconcentration only. Likewise, statistically significant increases in mean liver weights (p < 0.05)

were seen at the high-concentration group. Under the conditions of the study, a NOAEL and

LOAEL for maternal toxicity of 1 and 10 mg/m3, respectively, were indicated. 

No effects were observed on the maintenance of pregnancy or the incidence of resorptions. 

Mean fetal body weights were significantly decreased in the 25 mg/m3

 groups (p = 0.002) and in

the control group pair-fed 25 mg/m3 (p = 0.001). However, interpretation of the decreased fetal

body weight is difficult given the high incidence of mortality in the dams. Under the conditions

of the study, a NOAEL and LOAEL for developmental toxicity of 10 and 25 mg/m3

,

respectively, were indicated.

In trial two of the inhalation study, clinical signs of maternal toxicity seen at 10 and 25 mg/m3

were similar in type and incidence to those described for trial one. Maternal body weight gain

during treatment at 25 mg/m3 was less than controls, although the difference was not statistically

significant. In addition, 2 out of 12 dams died during treatment at 25 mg/m3. No other

treatment-related effects were reported, nor were any adverse effects noted for any of the

measurements of reproductive performance. Under the conditions of the study, a NOAEL and

LOAEL for maternal toxicity of 1 and 10 mg/m3, respectively, were indicated. 

Signs of developmental toxicity in this group consisted of statistically significant reductions in

pup body weight on Day 1 PP (6.1 g at 25 mg/m3 vs. 6.8 g in controls). On Days 4 and 22 PP,

pup body weights continued to remain lower than controls, although the difference was not

statistically significant (Day 4 PP: 9.7 g at 25 mg/m3 vs.10.3 in controls; Day 22 PP: 49.0 g at 25

mg/m3 vs. 50.1 in controls). No significant effects were reported following external examination

of the pups or with ophthalmoscopic examination of the eyes. Again, interpretation of these

effects is problematic given the high incidence of maternal mortality. Under the conditions of

the study, a NOAEL and LOAEL for developmental toxicity of 10 and 25 mg/m3

, respectively,

were indicated. 

In trial one of the oral study, three out of 25 dams died during treatment of 100 mg/kg APFO

during gestation (one death on GD 11; two on GD 12). Clinical signs of maternal toxicity in the

dams that died were similar to those seen with inhalation exposure. Food consumption and body

weights were reduced in treated animals compared to controls. No adverse signs of toxicity

were noted for any of the reproductive parameters such as maintenance of pregnancy or

incidence of resorptions. Likewise, no significant differences between treated and control

groups were noted for fetal weights, or in the incidences of malformations and variations; nor

were there any effects noted following microscopic examination of the eyes.

In trial two of the oral study, similar observations for clinical signs were noted for the dams as in

trial one. Likewise, no adverse effects on reproductive performance or in any of the fetal

observations were noted.
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64

3.8 Reproductive Toxicity Studies in Animals

An oral two-generation reproductive toxicity study of APFO in rats was conducted (York, 2002;

Butenhoff et al., 2004). Five groups of 30 Sprague-Dawley rats per sex per dose group were

administered APFO by gavage at doses of 0, 1, 3, 10, and 30 mg/kg-day six weeks prior to and

during mating. Treatment of the F0 male rats continued until mating was confirmed, and

treatment of the F0 female rats continued throughout gestation, parturition, and lactation. 

The F0 animals were examined twice daily for clinical signs, abortions, premature deliveries,

and deaths. Body weights of F0 male rats were recorded weekly during the dosage period and

then on the day of sacrifice. Body weights of F0 female rats were recorded weekly during the

pre- and cohabitation periods and then on gestation days (GD) 0, 7, 10, 14, 18, 21, and 25 (if

necessary) and on lactation days (LD) 1, 5, 8, 11, 15, and 22 (terminal body weight). Food

consumption values in F0 male rats were recorded weekly during the treatment period, while in

F0 female rats, values were recorded weekly during the precohabitation period, on GDs 0, 7, 10,

14, 18, 21, and 25 and on LDs 1, 5, 8, 11, and 15. 

Estrous cycling was evaluated daily by examination of vaginal cytology beginning 21 days

before the scheduled cohabitation period and continuing until confirmation of mating by the

presence of sperm in a vaginal smear or confirmation of a copulatory plug. On the day of

scheduled sacrifice, the stage of the estrous cycle was assessed. 

For mating, one male rat and one female rat per group were cohabitated for a maximum of 14

days. Female rats with evidence of sperm in a vaginal smear or copulatory plug were designated

as GD 0. Parental females were evaluated for length of gestation, fertility index, gestation index,

number and sex of offspring per litter, number of implantation sites, general condition of the

dam and litter during the postpartum period, litter size and viability, viability index, lactation

index, percent survival, and sex ratio. Maternal behavior of the dams was recorded on LDs 1, 5,

8, 15, and 22. 

F0 generation animals were sacrificed by carbon dioxide asphyxiation (day 106 to 110 of the

study for male rats, i.e., after completion of the cohabitation period; and LD 22 for female rats),

necropsied, and examined for gross lesions. Gross necropsy included examination of external

surfaces and orifices, as well as internal examination of tissues and organs. Individual organs

were weighed and organ-to-body weight and organ-to-brain weight ratios were calculated for the

brain, kidneys, spleen, ovaries, testes, thymus, liver, adrenal glands, pituitary, uterus with

oviducts and cervix, left epididymis (whole and cauda), right epididymis, prostate and seminal

vesicles, (with coagulating glands and with and without fluid). Tissues retained in neutral

buffered 10% formalin for possible histological evaluation included the pituitary, adrenal glands,

vagina, uterus, with oviducts, cervix and ovaries, right testis, seminal vesicles, right epididymis,

and prostate. Histological examination was performed on tissues from 10 randomly selected rats

per sex from the control and high dosage groups. All gross lesions were examined histologically. 

All F0 generation rats that died or appeared moribund were also examined.

Histological examination of the reproductive organs in the low- and mid-dose groups was

conducted in rats that exhibited reduced fertility by either failing to mate, conceive, sire, or

deliver healthy offspring; or for which estrous cyclicity or sperm number, motility, or

morphology were altered. Sperm number, motility, and morphology were evaluated in the left

cauda epididymis of F0 generation male rats; testicular spermatid concentrations were evaluated

in the left testis. The number and distribution of implantation sites were recorded in F0
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65

generation female rats. Rats that did not deliver a litter were sacrificed on GD 25 and examined

for pregnancy status. Uteri of apparently nonpregnant rats were examined to confirm the

absence of implantation sites. A gross necropsy of the thoracic, abdominal and pelvic viscera

was performed. Female rats without a confirmed mating date that did not deliver a litter were

sacrificed on an estimated day 25 of gestation. 

At scheduled sacrifice, after completion of the cohabitation period in F0 male rats and on LD 22

in F0 female rats, blood samples (10 males and 10 females each for the 10 and 30 mg/kg-day

dose groups; 3 males and 3 females for the control group) were collected for analysis of PFOA. 

The F1 generation pups in each litter were counted once daily. The litter sizes were not

standardized on day 4 or at anytime during lactation. Physical signs (including variations from

expected lactation behavior and gross external physical anomalies) were recorded for the pups

each day. Pup body weights were recorded on LDs 1, 5, 8, 15 and 22. On LD 12, all F1

generation male pups were examined for the presence of nipples. Pups that died before

examination of the litter for pup viability on LD 1 were evaluated for vital status at birth. Pups

found dead on LDs 2 to 22 were examined for gross lesions and for the cause of death. All F1

generation rats were weaned on LD 22 based on observed growth and viability of these pups. 

At weaning (LD 22), two F1 generation pups per sex per litter per group (60 male and 60 female

pups per group) were selected for continued evaluation, resulting in 600 total rats (300 rats per

sex) assigned to the five dosage groups. At least two male pups and two female pups per litter,

when possible, were selected. F1 generation pups not selected for continued observation for

sexual maturation were sacrificed. Three pups per sex per litter were examined for gross lesions. 

Necropsy included a single cross-section of the head at the level of the frontal-parietal suture and

examination of the cross-sectioned brain for apparent hydrocephaly. The brain, spleen and

thymus from one of the three selected pups per sex per litter were weighed and the brain, spleen,

and thymus from the three selected pups per sex per litter were retained for possible histological

evaluation. All remaining pups were discarded without further examination. 

 The F1 generation rats were given the same dosage level of the test substance and in the same

manner as their respective F0 generation sires and dams. Dosages were given once daily,

beginning at weaning and continuing until the day before sacrifice. F1 generation female rats

were examined for age of vaginal patency, beginning on day 28 postpartum (LD 28). F1

generation male rats were evaluated for age of preputial separation, beginning on day 39

postpartum (LD 39). Body weights were recorded when rats reached sexual maturation. 

Following sexual maturation, a table of random units was used to select one male and one female

per litter per group for continuation through mating to produce the F2 generation. The

remaining F1 animals were sacrificed.

Estrous cycling was evaluated daily by examination of vaginal cytology beginning 21 days

before the scheduled cohabitation period and continuing until confirmation of mating by the

presence of sperm in a vaginal smear or confirmation of a copulatory plug. On the day of

scheduled sacrifice, the stage of the estrous cycle was assessed. 

A table of random units was used to assign F1 generation rats to cohabitation, one male rat per

female rat. If random assignment to cohabitation resulted in the pairing of F1 generation

siblings, an alternate assignment was made. The cohabitation period consisted of a maximum of

14 days.
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66

Body weights of the F1 generation male rats were recorded weekly during the postweaning

period and on the day of sacrifice. Body weights of the F1 generation female rats were recorded

weekly during the postweaning period to cohabitation, and on GDs 0, 7, 10, 14, 18, 21 and 25 (if

necessary) and on LDs 1, 5, 8, 11, 15 and 22. Food consumption values for the F1 generation

male rats were recorded weekly during the dosage period. Food consumption values for the F1

generation female rats were recorded weekly during the postweaning period to cohabitation, on

GDs 0, 7, 10, 14, 18, 21 and 25 and on LDs 1, 5, 8, 11 and 15. Because pups begin to consume

maternal food on or about LD 15, food consumption values were not tabulated after LD 15. 

At scheduled sacrifice, the F1 animals were subjected to gross necropsy, and selected organs

were weighed and examined histologically as described above for the F0 animals. Sperm

analyses were also conducted as described for the F0 animals. 

F2 generation litters were examined after delivery to identify the number and sex of pups,

stillbirths, live births and gross alterations. Each litter was evaluated for viability at least twice

each day of the 22-day postpartum period. Dead pups observed at these times were removed

from the nesting box. Anogenital distance was measured for all live F2 generation pups on LDs

1 and 22, and F2 male pups were examined for the presence of nipples on LD 12. 

3.8.1 F0 Generation

3.8.1.1 F0 Males

One F0 male rat in the 30 mg/kg-day dose group was sacrificed on day 45 of the study due to

adverse clinical signs (emaciation, cold-to-touch, and decreased motor activity). Necroscopic

examination in that animal revealed a pale and tan liver, and red testes. All other F0 generation

male rats survived to scheduled sacrifice. Statistically significant increases in clinical signs were

also observed in male rats in the high-dose group that included dehydration, urine-stained

abdominal fur, and ungroomed coat. 

Significant reductions in body weight and body weight gain were reported for most of the dosage

period and continuing until termination of the study in the 3, 10, and 30 mg/kg-day dose groups. 

Absolute food consumption values were also significantly reduced during these periods in the 30

mg/kg-day dose group, while significant increases in relative food consumption values were

observed in the 3, 10, and 30 mg/kg-day within those same periods.

No treatment-related effects were reported at any dose level for any of the mating and fertility

parameters assessed, including numbers of days to inseminate, numbers of rats that mated,

fertility index, numbers of rats with confirmed mating dates during the first and second week of

cohabitation, and numbers of pregnant rats per rats in cohabitation. At necropsy, none of the

sperm parameters evaluated (sperm number, motility, or morphology) were affected by treatment

at any dose level. 

At necropsy, statistically significant reductions in terminal body weights were seen at 3, 10, and

30 mg/kg-day (6%, 11%, and 25% decrease from controls, respectively; p<0.05). Absolute

weights of the left and right epididymides, left cauda epididymis, seminal vesicles (with and

without fluid), prostate, pituitary, left and right adrenals, spleen, and thymus were also

statistically significantly reduced at 30 mg/kg-day. These absolute organ weight reductions are

probably due to reductions in body weight and not a reflection of target organ toxicity since the

organ-to-body weight ratios were either normal or significantly increased. The biological
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67

significance of the weight changes for the adrenals, however, is unclear since dose-related

histopathological changes were observed. The absolute weight of the seminal vesicles without

fluid was significantly reduced in the 10 mg/kg-day dose group. The absolute weight of the liver

was statistically significantly increased in all dose-groups. Absolute kidney weights were

statistically significantly increased in the 1, 3, and 10 mg/kg-day dose groups, but significantly

decreased in the 30 mg/kg-day group. Organ weight-to-terminal body weight ratios for the liver

and for the left and right kidney were statistically significantly increased in all treated groups. 

The biological significance of the weight changes observed in the liver and kidney is unknown

since no histopathology was conducted on these organs. Organ weight-to-terminal body weight

ratios for the brain were statistically significantly increased for the 3, 10, and 30 mg/kg-day dose

groups; organ weight-to-brain weight ratios were significantly reduced for some organs at the

high dose group, and significantly increased for other organs among all treated groups. 

No treatment-related effects were seen at necropsy or upon microscopic examination of the

reproductive organs, with the exception of increased thickness and prominence of the zona

glomerulosa and vacuolation of the cells of the adrenal cortex in 2/10 males in the 10 mg/kg-day

group and 7/10 males in the 30 mg/kg-day dose group. 

Serum analysis for the F0 generation males in the control, 10 and 30 mg/kg-day groups sampled

at the end of cohabitation showed that PFOA was present in all samples tested, including

controls. Control males had an average concentration of 0.0344+ 0.0148 :g/ml PFOA. Levels

of PFOA were similar in the two male dose groups; treated males had 51.1+9.30 and 45.3+12.6

:g/ml, respectively for the 10 and 30 mg/kg-day dose groups.

3.8.1.2 F0 Females

No treatment-related deaths or adverse clinical signs were reported in parental females at any

dose level. No treatment-related effects were reported for body weights, body weight gains, and

absolute and relative food consumption values.

There were no treatment-related effects on estrous cyclicity, mating or fertility parameters. 

None of the natural delivery and litter observations were affected by treatment, that is, the

numbers of dams delivering litters, the duration of gestation, the averages for implantation sites

per delivered litter, the gestation index (number of dams with one or more liveborn pups/number

of pregnant rats), the numbers of dams with stillborn pups, dams with all pups dying, liveborn

and stillborn pups viability index, pup sex ratios, and mean birth weights were comparable to

controls among all treated groups. 

Necropsy and histopathological evaluation were also unremarkable. Terminal body weights,

organ weights, and organ-to-terminal body weight ratios were comparable to control values for

all treated groups, except for kidney and liver weights. The absolute weights of the left and right

kidney, and the ratios of these organ weights-to-terminal body weight and of the left kidney

weight-to-brain weight were significantly reduced at the highest dose of 30 mg/kg-day. The

biological significance of these weight changes is not known since histopathology was not

conducted on the kidney. The ratio of liver weight-to-terminal body weight was significantly

reduced at 3 and 10 mg/kg-day, but there were no effects observed at 30 mg/kg-day or in

absolute liver weight. 

Serum PFOA levels were analyzed from the control, 10 and 30 mg/kg-day groups on LD 22. 

The samples were collected 24 hours after dosing. In the controls, serum PFOA was below the
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limits of quantitation (0.00528 :g/ml). Levels of PFOA found in female sera increased between

the two dose groups; treated females had an average concentration of 0.37+0.0805 and

1.02+0.425 :g/ml, respectively for the 10 and 30 mg/kg-day dose groups. 

3.8.2 F1 Generation

No effects were reported at any dose level for the viability and lactation indices. No differences

between treated and control groups were noted for the numbers of pups surviving per litter, the

percentage of male pups, litter size and average pup body weight per litter at birth. At 30 mg/kgday, one pup from one dam died prior to weaning on lactation day 1 (LD1). Additionally, on

lactation days 6 and 8, statistically significant increases in the numbers of pups found dead were

observed at 3 and 30 mg/kg-day. According to the study authors, this was not considered to be

treatment related because they did not occur in a dose-related manner and did not appear to

affect any other measures of pup viability including numbers of surviving pups per litter and live

litter size at weighing. An independent statistical analysis was conducted by US EPA (2002b). 

No significant differences were observed between dose groups and the response did not have any

trend in dose. 

The authors did not present the mean pup body weights for the male and female pups separately.

Pup body weight on a per litter basis (sexes combined) was reduced throughout lactation in the

30 mg/kg-day group, and statistical significance (p < 0.01) was achieved on days 1 (9.5%), 5

(9.6%), and 8 (10.5%). Of the pups necropsied at weaning, no statistically significant,

treatment-related differences were observed for the weights of the brain, spleen and thymus and

the ratios of these organ weights to the terminal body weight and brain weight. 

3.8.2.1 F1 Males

Significant increases in treatment-related deaths (5/60 animals) were reported in F1 males in the

high dose group of 30 mg/kg-day between days 2-4 postweaning. One rat was moribund

sacrificed on day 39 postweaning and another was found dead on day 107 postweaning. 

Statistically significant increases in clinical signs of toxicity were also observed in F1 males

during most of the entire postweaning period. These signs included an increased incidence of

annular constriction of the tail at all doses, with statistical significance at the 1, 10, and 30

mg/kg-day; a significant increase at 10 and 30 mg/kg-day in the number of male rats that were

emaciated; and a significant increase in the incidence of urine-stained abdominal fur, decreased

motor activity, and abdominal distention at 30 mg/kg-day. 

Statistically significant reductions in body weight gain were observed at 10 and 30 mg/kg-day

during days 8-15, 22-29, 29-36, 43-50, and 50-57 postweaning. Body weight gains were also

significantly reduced in the 30 mg/kg-day group on days 1-8, 15-22, 36-43, 57-64, and 64-70

postweaning. Body weights were significantly reduced in the 10 mg/kg-day group beginning on

postweaning day 36, and in the 30 mg/kg-day group beginning on postweaning day 8. In the 3

mg/kg-day group, mean body weight gain was significantly reduced on days 43-50 and 57-64

postweaning, and mean body weights were significantly reduced on days 106 and 113

postweaning. In the 1 mg/kg-day group, mean body weight gain was significantly reduced on

days 15-22 and 43-50 postweaning, and mean body weights were significantly reduced on days

50, 57, 64, 70, 99, 106 and 113 postweaning. For all groups, there was a significant, doserelated reduction in mean body weight gain for the entire dosing period (days 1-113). Absolute

food consumption values were significantly reduced at 10 and 30 mg/kg-day during the entire
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precohabitation period (days 1-70 postweaning), while relative food consumption values were

significantly increased.

Statistically significant (p< 0.01) delays in sexual maturation (the average day of preputial

separation) were observed in high-dose animals versus concurrent controls (52.2 days of age

versus 48.5 days of age, respectively).

No apparent effects were observed on any of the mating or fertility parameters including fertility

and pregnancy indices (number of pregnancies per number of rats that mated and rats in

cohabitation, respectively), the number of days to inseminate, the number of rats that mated, and

the number of rats with confirmed mating dates during the first week. No statistically

significant, treatment-related effects were observed on any of the sperm parameters (motility,

concentration, or morphology).

Necroscopic examination revealed statistically significant treatment-related effects at 3, 10, and

30 mg/kg-day ranging from tan areas in the lateral and median lobes of the liver to moderate to

slight dilation of the pelvis of one or both kidneys.

Statistically significant, dose-related decreases in terminal body weights were observed in the 1,

3, 10, and 30 mg/kg-day dose groups(6%, 6%, 11%, and 22% decrease from controls,

respectively; p<0.01 at 1 and 3 mg/kg-day, p<0.05 at 10 and 30 mg/kg-day). The absolute and

relative weights of the liver were statistically significantly increased in all treated groups (p <

0.01) and was accompanied by histopathological changes. The absolute weights of the left

and/or right kidneys were statistically significantly increased in the 1 and 3 mg/kg-day dose

groups and statistically significantly decreased in the 30 mg/kg-day dose group. Organ weightto-terminal body weight and brain weight ratios for the kidney were statistically significantly

increased in all treated groups. The biological significance of the effects on kidney weight is

unknown since histopathology was not conducted on that organ. All other organ weight changes

observed (thymus, spleen, left adrenal, brain, prostate, seminal vesicles, testes, and epididymis)

are probably due to decrements in body weight and not a reflection of target organ toxicity since

the absolute weights of these organs was significantly reduced while the relative weights were

either normal or significantly increased. However, the biological significance of the weight

changes observed in the adrenal is unclear since histopathological changes were also observed. 

Histopathologic examination of the reproductive organs was unremarkable; however, treatmentrelated microscopic changes were observed in the adrenal glands of high-dose animals

(cytoplasmic hypertrophy and vacuolation of the cells of the adrenal cortex) and in the liver of

animals treated with 3, 10, and 30 mg/kg-day (hepatocellular hypertrophy). No other treatmentrelated effects were reported. 

3.8.2.2 F1 Females

A statistically significant increase in treatment-related mortality (6/60 animals) was observed in

F1 females on postweaning days 2-8 at the highest dose of 30 mg/kg-day. No adverse clinical

signs of treatment-related toxicity were reported for any dose level during any time of the study

period. 

Statistically significant decreases in body weights were observed in high-dose animals on days 8,

15, 22, 29, 50, and 57 postweaning, during precohabitation (recorded on the day cohabitation

began, when F1 generation rats were 92-106 days of age), and during gestation and lactation.
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Body weight gain was significantly reduced during days 1-8 and 8-15 postweaning. Statistically

significant decreases in absolute food consumption were observed during days 1-8, 8-15, and 15-

22 postweaning, during precohabitation and during gestation and lactation in animals treated

with 30 mg/kg-day. Relative food consumption values were comparable across all treated

groups.

Statistically significant (p< 0.01) delays in sexual maturation (the average day of vaginal

patency) were observed in high-dose animals versus concurrent controls (36.6 days of age versus

34.9 days of age, respectively).

Prior to mating, the study authors noted a statistically significant increase in the average numbers

of estrous stages per 21 days in high-dose animals (5.4 versus 4.7 in controls). For this

calculation, the number of independent occurrences of estrus in the 21 days of observation was

determined. This type of calculation can be used as a screen for effects on the estrous cycle, but

a more detailed analysis should then be conducted to determine whether there is truly an effect. 

3M Company (2002) recently completed an analysis that showed there were no effects on the

estrous cycle; there were no differences in the number of females with > 3 days of estrus or with

> 4 days of diestrus in the control and high dose groups. Analyses conducted by the US EPA

(2002a) also demonstrated that there were no differences in the estrous cycle among the control

and high dose groups. The cycles were evaluated as having either regular 4-5 day cycles, uneven

cycling (defined as brief periods with irregular pattern) or periods of prolonged diestrus (defined

as 4-6 day diestrus periods) extended estrus (defined as 3 or 4 days of cornified smears), possibly

pseudopregnant, (defined as 6-greater days of leukocytes) or persistent estrus (defined as 5-or

greater days of cornified smears). The two groups were not different in any of the parameters

measured. Thus, the increase in the number of estrous stages per 21 days that was noted by the

study authors is due to the way in which the calculation was done, and is not biologically

meaningful.

No effects on any of the mating and fertility parameters (numbers of days in cohabitation,

numbers of rats that mated, fertility index, rats with confirmed mating dates during the first week

of cohabitation and number of rats pregnant per rats in cohabitation). 

All natural delivery observations were unaffected by treatment at any dose level. Numbers of

dams delivering litters, the duration of gestation, averages for implantation sites per delivered

litter, the gestation index (number of dams with one or more liveborn pups/number of pregnant

rats), the numbers of dams with stillborn pups, dams with all pups dying and liveborn and

stillborn pups were comparable among treated and control groups.

The terminal body weights and absolute and relative pituitary weights are shown in Table 19. 

No treatment-related effects were observed in the terminal body weights of the F1 females. The

absolute weight of the pituitary, the pituitary weight-to-terminal body weight ratio, and the

pituitary weight-to-brain weight ratio were statistically significantly decreased at 3 mg/kg-day

and higher. Since there is not a clear dose-response relationship and histologic examination did

not reveal any lesions, the biological significance of the pituitary weight data is problematic. No

other differences were reported for the absolute weights or ratios for other organs evaluated. No

treatment-related effects were reported following macroscopic and histopathologic examinations. 
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Table 19

Summary of Body and Pituitary Weights of F1 Females

Group

Terminal

Body Weight (g)a PituitaryWeight (g)a Pituitary/Body Weight (%)a Pituitary/BrainWeight (%)a

Control 322.9 + 23.4 0.017 + 0.004 5.46 + 1.6 0.84 + 0.21

1 mg/kg-day 321.7 + 24.2 0.016 + 0.003 5.0 + 0.86 0.76 + 0.14

3 mg/kg-day 329.2 + 21.5 0.015 + 0.003* 4.65 + 0.86* 0.74 + 0.14*

10 mg/kgday325.1 + 23.5 0.015 + 0.002* 4.72 + 0.90* 0.72 + 0.12*

30 mg/kgday315.7 + 20.9 0.015 + 0.003* 4.75 + 1.03* 0.72 + 0.16**

a- Mean + SD

* Significantly different from control at p < 0.05

** Significantly different from control at p < 0.01

3.8.3 F2 Generation

No treatment-related adverse clinical signs were observed at any dose level. Dead or stillborn

pups were noted in both the control and treated groups. The deaths occurred on lactation days 1-

18 with the majority occurring on days 1-6. However, there was no dose-response relationship

and therefore were unlikely related to treatment. Statistically significant increases (p< 0.01) in

the number of pups found dead were observed on lactation day 1 in the 3 and 10 mg/kg-day

groups. According to the study authors, this was not considered to be treatment related because

they did not occur in a dose-related manner and did not appear to affect any other measures of

pup viability including numbers of surviving pups per litter and live litter size at weighing. An

independent statistical analysis was conducted by US EPA (2002b). No significant differences

were observed between dose groups and the response did not have any trend in dose. 

No effects were reported at any dose level for the viability and lactation indices. No differences

between treated and control groups were noted for the numbers of pups surviving per litter, the

percentage of male pups, litter size and average pup body weight per litter when measured on

LDs 1, 5, 8, 15, or 22. Anogenital distances measured for F2 male and female pups on LDs 1

and 22 were also comparable among the five dosage groups and did not differ significantly. 

Terminal body weights in F2 pups were not significantly different from controls. Absolute

weights of the brain, spleen and thymus and the ratios of these organ weights-to-terminal body

weight and to brain weight were also comparable among treated and control groups. There were

no treatment-related effects following necroscopic examination, with the exception of no milk in

the stomach of the pups that were found dead. 

3.8.4 Conclusions

Dosing with APFO at 30 mg/kg-day resulted in a delay in the onset of sexual maturation in both

male and female F1 offspring. The authors of the study contend that the delays in sexual

maturation (preputial separation or vaginal patency) observed in high-dose animals are due to the
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72

fact that these animals have a decreased gestational age, a variable which they have defined as

the time in days from evidence of mating in the F0 generation until evidence of sexual

maturation in the F1 generation. The authors state that gestational age appeared to be decreased

in high-dose animals at the time of acquisition (the time when sexual maturation was reached),

which they believe meant the animals in that group were younger and more immature than the

control group, in which there was no significant difference in sexual maturation. 

In order to test this hypothesis, the authors covaried separately the decreases in body weight and

in gestational age with the delays in sexual maturation in order to determine whether or not body

weights and gestational age were a contributing factor. When the body weight was covaried

with the time to sexual maturation, the time to sexual maturation showed a dose related delay

that was statistically significant at p< 0.05. This suggests that the delay in sexual maturation was

partly related to body weight, but not entirely. When gestational age was covaried with the time

to sexual maturation, there was no significant difference in the time of onset of sexual

maturation between controls and high-dose animals. This indicates that the effect of delayed

sexual maturation could possibly be attributed to decreased gestational age.

While it is known and commonly accepted that changes in the body weights of offspring can

affect the time to sexual maturation, whether or not gestational age, as defined by the authors,

also affects the time of sexual maturation is purely speculative, especially since there were no

data provided by the authors to support this relationship. Additionally, covaring gestational age

with time to sexual maturation is problematic from a statistical standpoint. Since there was no

significant change in the length of gestation at 30 mg/kg-day, based on the authors’ definition of

‘gestational age’, the decreases in gestational age would have to be due mostly to changes in

time to sexual maturation. Therefore, sexual maturation is essentially being covaried with itself. 

Still, even if a relationship between gestational age and time to sexual maturation were shown, it

merely offers an explanation for the observed delays in sexual maturation in high-dose animals,

but does not diminish its significance. 

A variety of endpoints are evaluated throughout different lifestages in a two generation

reproductive toxicity study. Therefore, some of the endpoints may be indicative of

developmental/reproductive toxicity, while others may be indicative of adult toxicity. The

selection of developmental endpoints (and the appropriate LOAELs and NOAELs) for this study

was based on the Agency’s Developmental Toxicity Risk Assessment Guidelines (EPA, 1991)

and the Agency’s Reproductive Toxicity Risk Assessment Guidelines (EPA, 1996). According

to the guidelines, the period of exposure for developmental toxicity is prior to conception in

either parent, through prenatal development and continuing until sexual maturation. In contrast,

the period during which a developmental effect may be manifested includes the entire lifespan of

the organism. For selection of the developmental endpoints from the two generation

reproductive toxicity study, attention was focused on effects that were noted during the period of

developmental exposure. Thus, only effects that occurred up to sexual maturation were

considered relevant for assessing developmental toxicity. Effects occurring after sexual

maturation were considered relevant for assessing adult toxicity since it was not possible to

determine whether the effects were due to developmental and/or adult exposures.

Therefore, under the conditions of the study, the LOAEL for F0 parental males is 1 mg/kg-day,

the lowest dose tested, based on significant increases in absolute and relative liver weight. A

NOAEL for the F0 parental males could not be determined since the increases in liver weight

were seen at all doses tested. The NOAEL for F0 parental females is 30 mg/kg-day, the highest

dose tested.
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A variety of developmental/reproductive effects were noted in the F1 generation. During

lactation, there was a significant reduction in F1 mean body weight on a litter basis (sexes

combined) in the 30 mg/kg-day group. F1 males in the 10 and 30 mg/kg-day groups exhibited a

significant reduction in body weight gain during days 8-50 postweaning, and body weights were

significantly reduced in the 10 mg/kg-day group beginning on postweaning day 36, and in the 30

mg/kg-day beginning on postweaning day 8. F1 females in the 30 mg/kg-day group exhibited a

significant reduction in body weight gain on days 1-15 postweaning, and in body weights

beginning on day 8 postweaning. There was a significant increase in mortality mainly during the

first few days after weaning, and a significant delay in the timing of sexual maturation for F1

males and females in the 30 mg/kg-day group. For F1 males, the LOAEL for

developmental/reproductive toxicity was 10 mg/kg-day, and the NOAEL was 3 mg/kg-day. For

F1 females, the LOAEL for developmental/reproductive toxicity was considered to be 30 mg/kgday, and the NOAEL was 10 mg/kg-day. 

The LOAEL for adult systemic toxicity in the F1 males is 1 mg/kg-day based on significant,

dose-related decreases in body weights and body weight gains (observed prior to and during

cohabitation and during the entire dosing period), and in terminal body weights; and significant

changes in absolute and relative liver weights. A NOAEL for the F1 males could not be

determined since these effects were seen at all doses tested. 

The NOAEL and LOAEL for adult systemic toxicity in the F1 females are 10 and 30 mg/kg-day,

respectively, based on statistically significant decreases in body weight and body weight gains. 

The NOAEL for developmental/reproductive toxicity in the F2 offspring was 30 mg/kg-day. No

treatment-related effects were observed at any doses tested in the study. However, it should be

noted that the F2 pups were sacrificed at weaning, and thus it was not possible to ascertain the

potential post-weaning effects that were noted in the F1 generation.

3.9 Mode of Action and Summary of Weight of Evidence

3.9.1 Epidemiology Studies

All of the epidemiologic data available on PFOA are based on occupational studies, most of

which were routine biomonitoring efforts conducted by 3M Company. With the exception of

one, all of the studies were cross-sectional and mostly analyzed males, the majority of workers at

these plants. Consequently, reproductive and developmental outcomes have not been studied. 

In the 3M mortality studies, the only statistically significant association reported was for prostate

cancer mortality and employment duration in a plant that manufactures PFOA; however, this

association was not observed in a follow-up study which included more specific exposure

measures (Gilliland and Mandel, 1993; Alexander, 2001a). In the Dupont cancer study, bladder

and kidney cancer incidence was elevated among employees (Dupont, 2003). However, very

little other data on exposures, including other chemicals at the plant, were available. A study on

hormone levels in male workers indicated an increase in estradiol levels in workers with the

highest PFOA serum levels; however, the results may have been confounded by body mass

index (Olsen, et al., 1998a). There were no changes of note in other hormone levels in these

workers. Another study of CCK levels in employees, in which increased levels have been linked

to pancreas acinar cell adenomas in rats, did not report increases in workers (Olsen, et al., 1998b;

Olsen et al., 2000). In addition, cholesterol and triglyceride levels in workers were positively

associated with PFOA exposures, inconsistent with the hypolipidemic results reported in rat
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studies. There was also a statistically significant positive association reported for PFOA and T3

but not any other thyroid hormones (Olsen, et al., 2001e). 

Overall, there were no notable health effects reported in fluorochemical workers which can be

directly attributed to PFOA exposure. However, there were many limitations in these studies

(detailed in Section 3.1) which need to be considered.

3.9.2 Metabolism and Pharmacokinetics

Little information is available concerning the pharmacokinetics of PFOA and its salts in humans. 

An ongoing 5-year, half-life study in 7 male and 2 female retired workers has suggested a mean

serum PFOA half-life of 4.37 years (range, 1.50 – 13.49 years). Metabolism and

pharmacokinetic studies in non-human primates are limited to a study of 3 male and 3 female

cynomolgus monkey administered a single IV dose of 10 mg/kg potassium PFOA. In male

monkeys, the average serum half life was 20.9 days. In female monkeys, the average serum half

life was 32.6 days.

Studies in adult rats have shown that PFOA is absorbed following oral, inhalation, and dermal

exposures. Serum pharmacokinetic parameters of PFOA have been evaluated in adult SpragueDawley rats following gavage administration, and in Wistar rats following IV administration. 

The distribution of PFOA has been examined in tissues of adult rats following administration by

gavage and by i.v. and i.p. injection. PFOA distributes primarily to the liver, serum, and kidney,

and to a lesser extent, other tissues of the body. It does not partition to the lipid fraction or

adipose tissue. Although investigated in several studies, there is no evidence that PFOA is

metabolized. In adult rats, there is evidence of enterohepatic circulation of PFOA. The urine is

the major route of excretion of PFOA in the female rat, while the urine and feces are both main

routes of excretion in male rats. There are gender differences in the elimination of PFOA in rats. 

In female rats, following oral administration, estimates of the serum half-life were dependent on

dose and ranged from approximately 2.8 to 16 hours, while in male rats estimates of the serum

half-life following oral administration were independent of dose and ranged from approximately

138 to 202 hours. In female rats, elimination of PFOA appears to be biphasic with a fast phase

and a slow phase. The rapid excretion of PFOA by female rats is believed to be due to active

renal tubular secretion (organic acid transport system); this renal tubular secretion is believed to

be hormonally controlled. Hormonal changes during pregnancy do not appear to cause a change

in the rate of elimination in rats. 

No studies have been conducted to specifically examine the absorption, metabolism or

elimination of PFOA in the developing rat. However, recent studies have been conducted to

examine the concentrations of PFOA in the developing Sprague-Dawley rat, and to determine

when the gender difference in elimination of PFOA becomes apparent. In addition, several

studies have examined the serum and tissue distribution of PFOA in newly weaned Wistar rats. 

These studies have shown that PFOA readily crosses the placenta and is present in the breast

milk of rats. During lactation and for the first several weeks after weaning, the elimination of

PFOA is similar in males and females. Between 4-5 weeks of age, the elimination in male rats

assumes the adult pattern and the gender difference becomes readily apparent. Distribution

studies in the postweaning rat have shown that PFOA is distributed primarily to the serum, liver,

and kidney.

It has been suggested that PFOA is circulated around the body by noncovalently binding to

plasma proteins. Several studies have investigated the binding of PFOA to plasma proteins of
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rats, humans or monkeys to gain understanding of its absorption, distribution and elimination

and species and gender differences. There is limited additional information on the metabolism

and pharmacokinetics of PFOA in mice, rabbits and dogs. 

3.9.3 Mode of Action Analyses and Cancer Descriptor

There has been a great deal of scientific debate and research into rodent liver toxicity and liver

tumors that are induced through the activation of PPAR", and several groups of scientists have

summarized the state of the science over the years (IARC, 1995; Cattley et al. 1998). More

recently, the ILSI Risk Science Institute convened a workgroup to re-examine the mode of action

of PPAR" agonist-induced rodent liver tumors as well as to evaluate the mode(s) of action for

Leydig cell and pancreatic acinar cell tumors, which also are observed frequently in rats with

PPAR" agonists (Klaunig et al., 2003). This effort utilized and extended guidance that the US

EPA (1999) and IPCS (Sonich-Mullin et al., 2001) have provided for assessing information for

the delineation of an animal mode of action (MOA) for specific tumor types. This guidance is

also applicable for assessing MOAs for non-tumor responses. In this context, it is important to

note that the term “mode of action” is defined as the sequence of key cellular and biochemical

events (measurable parameters) that result in a toxicological effect. The analysis does not

require knowledge of “mechanism of action” which implies a more detailed understanding of the

molecular basis of the toxicological effect. 

Recently, EPA’s Office of Prevention, Pesticides and Toxic Substances (OPPTS) presented a

draft of a proposed OPPTS science policy on PPAR" mediated hepatotoxicity and

heptacarcinogenesis to the FIFRA Science Advisory Panel (SAP) in December, 2003 (OPPTS,

2003). The OPPTS guidance document describes the approach the Office proposes to use to

evaluate the scientific information regarding the mode of action of PPAR" agonists in rodent

hepatocarcinogenesis and the relevance of this mode of action for human hepatocarcinogenesis. 

Other tumor types (e.g. Leydig cell and pancreatic acinar cell tumors) that may be associated

with PPAR"-agonists are also briefly described. The document provides an overview of the

evidence for a PPAR"-agonist mode of action for liver tumors in rodents, and an overview of all

known age and species differences in the key events. The document also provides proposed

guidance on the data needed to demonstrate that rodent liver toxicity and tumors have arisen as a

result of a PPAR" agonist mode of action, and the relevance of this mode of action for humans. 

As described in section 3.5, several studies of PFOA in adult rats have shown that the liver is a

principle target organ, and long-term exposure results in liver adenomas, Leydig cell adenomas,

and pancreatic acinar cell tumors. The mode of action information for each of these effects are

described below and are described in the context of the work of Klaunig et al. (2003), the

proposed OPPTS science policy and the response from the FIFRA SAP (FIFRA SAP, 2004).

3.9.3.1 Mode of Action Analysis of Liver Toxicity and Liver Adenomas in Rats

There are a number of possible modes of action for hepatocarcinogenesis of chemicals. As

summarized in Section 3.4, the weight of evidence from short-term genotoxicity assays suggest

that PFOA is not a DNA-reactive compound. 

As mitochondria play a major role in cell signaling and apoptotic modes of cell death, and several

structurally related perfluorinated compounds have been shown to manifest their toxicity by

interfering with mitochondria biogenesis and bioenergetics, the effects of PFOA on

mitochondrial biogenesis and bioenergetics were investigated. A number of studies have shown
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that mitochondria biogenesis in liver was increased following treatment of rats with PFOA (e.g.,

Berthiaume and Wallace, 2002). PFOA has also been demonstrated to uncouple oxidative

phosphorylation in mitochondria of the liver from rats exposed to PFOA in the diet (Keller et al.,

1992). At high concentrations, PFOA caused a small increase in resting respiration rate and

slightly decreases the membrane potential. The observed effects are believed to be attributed to a

slight increase in nonselective permeability of the mitochondria membranes caused by the

surface-active property of the compound (Starkov and Wallace, 2002). Treatment with 200 :M

PFOA was found to cause a dramatic increase in the cellular content of reactive oxygen species

(ROS) in human hepG2 cells. The activation of caspase-9 and apoptosis by PFOA observed was

postulated to be the result of the disruption of mitochondria membrane and accumulation of ROS

(Panaretakis et al., 2001). Further research is needed, however, to elucidate how apoptosis is

involved in tumorigenesis of PFOA. 

Gap junctional intercellular communication (GJIC), a process by which cells exchange ions,

second messages, and other small molecules, is important for normal growth, development, and

differentiation, as well as maintenance of homeostasis in muticellular organisms. Because tumor

formation requires loss of homeostasis and many tumor promoters inhibit GJIC, it has been

hypothesized that GJIC may play a role in carcinogenesis (Trosko et al., 1998). PFOA has been

demonstrated to inhibit GJIC in liver cells in vitro and in vivo (Upham et al., 1998). Since

inhibition of GJIC is a widespread phenomenon, and the effect by PFOA was neither species nor

tissue specific and was reversible, the significance of GJIC inhibition in regard to the mode of

carcinogenic action of PFOA is unknown. 

Estrogen has been shown to promote hepatocarcinogenesis in rats (Yager and Yager, 1980;

Cameron et al., 1982). However, more research is needed to support the involvement of this

MOA in the hepatocarcinogenesis of PFOA. 

Low and sporadic incidences of liver necrosis were noted in both control and treated rats in the

subchronic and chronic toxicity studies of PFOA. Liver necrosis and regenerative cell

proliferation may play a role in promoting pre-initiated cells in carcinogenesis. The involvement

of necrosis in the liver tumor induction by PFOA in rats remains to be investigated. 

While available data are not sufficient to support any of the above MOAs for the liver tumor

induction by PFOA, there is strong evidence to conclude that the liver toxicity and liver

adenomas that are observed in rats following exposure to PFOA result from a PPAR"-agonist

MOA. As described in Klaunig et al. (2003) and OPPTS (2003), the MOA of PPAR"-agonist

induced liver toxicity and liver tumors involves four causal key events which are shown in Figure

1. (Page 78, below.) The first key event is activation of PPAR" (which regulates the

transcription of genes involved in peroxisome proliferation, cell cycle control, apoptosis, and

lipid metabolism). Activation of PPAR" leads to an increase in cell proliferation and a decrease

in apoptosis, which in turn leads to preneoplastic cells and further clonal expansion and formation

of liver tumors. Of these key events, only PPAR" activation is highly specific for this MOA

while cell proliferation/apoptosis and clonal expansion are common to other modes of action. As

depicted in Figure 1, there are also several “associative” events that are markers of PPAR"

agonism but are not directly involved in the etiology of liver tumors. These include peroxisome

proliferation (a highly specific indicator that this MOA is operative) and peroxisomal gene

expression. Peroxisomal proliferation may also result in hepatocyte oxidative stress which may

contribute to the mode of action by causing indirect DNA damage and leading to mutations, or by

stimulating cell proliferation. However, increases in oxidative damage to DNA have not been

unambiguously demonstrated for PPAR" agonists. Oxidative stress is a general phenomenon,
SAB Review Draft; Do Not Cite or Quote

77

and thus does not represent a highly specific marker for PPAR"-agonist induced liver

carcinogenesis.

OPPTS (2003) provided some guidance on information that would help establish that a chemical

is inducing liver toxicity and tumors via a PPAR" agonist MOA. This includes in vitro evidence

of PPAR" agonism (i.e., evidence from an in vitro receptor assay), in vivo evidence of an increase

in number and size of peroxisomes, increases in the activity of acyl CoA oxidase, and hepatic cell

proliferation. The in vivo evidence should be collected from studies designed to provide the data

needed to show dose-response and temporal concordance between precursor events and liver

tumor formation. Other information that is desirable and may strengthen the weight of evidence

for demonstrating that a PPAR" agonist MOA is operative includes data on hepatic CYP4A1

induction, palmitoyl CoA activity, hepatocyte hypertrophy, increase in liver weights, decrease in

the incidence of apoptosis, increase in microsomal fatty acid oxidation, and enhanced formation

of hydrogen peroxide.

There is sufficient information to demonstrate the key events for a PPAR" agonist MOA

following exposure to PFOA in rodents. It has been well documented that PFOA is a potent

peroxisome proliferator, inducing peroxisome proliferation in the liver of rats and mice (e.g.,

Ikeda et al., 1985; Pastoor et al., 1987; Sohlenius et al., 1992). Maloney and Waxman (1999)

demonstrated that PFOA activates PPARα using COS1 cells transfected with a luciferase reporter

gene. Maximal transcriptional activity with PFOA was seen at 10 μM in the mouse PPARα (and

at 20 μM in human PPARα). Like many other peroxisome proliferators, PFOA has also been

shown to cause hepatomegaly (an early biomarker of peroxisome proliferator

hepatocarcinogenesis) in rats (Takagi et al., 1992; Cook, 1994) and mice (Kennedy, 1986), and

induce oxidative DNA damage in liver of rats (Takagi et al., 1991).
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78

Key Events in the Mode of Action for PPARαAgonist Induced Rodent Liver Tumors

Causative Events

PPARα Agonist

Activation of PPAR α

Cell Proliferation

Decreased Apoptosis

Preneoplastic Foci

Clonal Expansion

Liver Tumors

•Expression of Peroxisomal Genes

•Increase in Peroxisomes (number 

& size)

Associative Events*

*Although there are other biological events (e.g., Kupffer cell mediated events, inhibition of gap junctions), the 

measurements of peroxisome proliferation and peroxisomal enzyme activity (in particular acyl-CoA) are widely used as 

reliable markers of PPARα activation.

Figure 1

From OPPTS (2003)
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79

Several studies have been conducted to examine the dose-response and temporal relationships

among key endpoints. The temporal and dose-response relationship of measures of peroxisome

proliferation, hepatocellular hypertrophy, liver weight, and liver histopathology have been

examined in male Sprague-Dawley rats following 4, 7 and 13 weeks of administration of dietary

PFOA at doses ranging from 0 - 6.5 mg/kg-day (Palazzolo, 1993). The results are summarized in

Table 20. There was no evidence of peroxisome proliferation, hepatocellular hypertrophy or liver

weight increases at 0.06 mg/kg-day. At doses ranging from 0.64 to 6.5 mg/kg-day there is a clear

relationship between peroxisome proliferation (indicated by increased palmitoyl CoA oxidase

activity), hepatocellular hypertrophy and increases in liver weight at all time points. There was

no evidence of hepatocellular necrosis.

Table 20 

Summary of Liver Effects in Male Sprague-Dawley Rats Fed APFO for 90 Days

Parameter Week Dose (mg/kg-day)1

0 a 0 b 0.06 0.64 1.94 6.5

Palmitoyl CoA 

Oxidase (IU/G)

4

7

13

8 (0.5)

7 (1.5)

8 (0.9)

5 (0.4)

7 (1.5)

5 (1.1)

9 (1.7)

7 (0.8)

8 (1.9)

14 (3.8)c

18 (5.5)

10 (2.1)

24 (11.4)c

32 (12.2)c

14 (3.4)c

37 (14.8)cd

54 (35.3)cf

17 (4.5)cd

Hepatocellular 

Hypertrophy

4

7

13

0/15

0/15

0/15

0/15

0/15

0/15

0/15

0/15

0/15

12/15

12/15

13/15

15/15

15/15

14/15

14/15

15/15

15/15

Hepatocellular 

Necrosis, 

Coagulative

4

7

13

0/15

0/15

0/15

1/15

0/15

0/15

0/15

0/15

1/15

0/15

0/15

0/15

1/15

0/15

1/15

2/15

1/15

0/15

Absolute Liver 

Weight (g)

4

7

13

16.34

(2.14)

17.78

(2.12)

19.73

(2.01)

15.83

(1.13)

16.91

(2.22)

16.30

(1.62)

15.45

 (1.71)

17.68

(??)

18.03

(2.81)

17.89 

(2.13)

19.42

(2.10)

20.44

(2.87)

23.23 

(2.83)c

27.76

(3.51)c

22.74

(4.21)

25.44 

(1.89)cd

27.76

(3.51)cd

26.78

(5.47)cd

Liver/Body 

Weight (%)

4

7

13

3.97 

(0.37)

3.75

(0.29)

3.53

(0.28)

4.07 

(0.27)

3.76

(0.37)

3.23

(0.23)

3.73 

(0.23)

3.64

(0.33)

3.24

(0.30)c

4.48 

(0.32)d

4.12

(0.37)

3.69

(0.32)

5.77 

(0.60)d

5.14

(0.53)c

4.21

(0.56)c

6.73 

(0.49)de

6.06

(0.59)cd

5.49

(0.84)cd

1- Mean (SD)

a - non-pair-fed controls

b - pair-fed controls

c - statistically significant at p < 0.05 with the non-pair-fed control

d - statistically significant at p < 0.05 with the pair-fed control

e - calculated using the non-pair-fed control

f - calculated using the pair-fed control
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80

Liu et al. (1996) characterized the dose-response relationships of several key endpoints in male

CD rats exposed to doses of 0.2, 2, 20, and 40 mg/kg-day PFOA for 14 days. These endpoints

included liver weight, hepatic β-oxidation, hepatic aromatase (P450 19A1), and hepatic total

cytochrome P450. The NOEL for these endpoints was found to be 0.2 mg/kg-day with significant

changes observed at > 2 mg/kg-day for all endpoints. Thus, the studies of Palazzolo (1993) and

Liu et al. (1996) demonstrate evidence of peroxisome proliferation at doses close to and below

doses that result in liver adenomas following chronic exposures (1.5 and 15 mg/kg-day).

Biegel et al. (2001) examined the temporal relationship between relative liver weights, hepatic βoxidation, and hepatic cell proliferation, and hepatic adenomas were evaluated in CD rats

following PFOA exposure for 1, 3, 6, 9, 12, 15, 18, 21, and 24 months. Relative liver weights and

hepatic β-oxidation were increased at all time points. Hepatic cell proliferation was numerically

increased relative to the pair-fed control at 9, 15, 18, and 21 months. The liver endpoints (weight,

β-oxidation, and cell proliferation) were all elevated well before the first occurrence of liver

adenomas, which occurred after 12 months of treatment.

In summary, these data clearly demonstrate that PFOA induces liver toxicity and adenomas via a

PPAR" agonist MOA in rats. PFOA activates the PPAR" and the requisite dose-response and/or

temporal associations of the key events for the PPAR" mode of action with the liver adenomas

have been characterized. 

3.9.3.2 Human Relevance of the Rat PPAR"-agonist Induced Liver Toxicity and Liver

Adenomas

There has been substantial scientific interest regarding the role of peroxisome proliferation in

rodent hepatocarcinogenesis and its relevance for human carcinogenesis. Several scientific

groups have examined the state of the science on PPAR" agonist-induced rodent liver tumors

over the years. In 1995, a workgroup convened under the auspices of the International Agency

for Research on Cancer concluded that the MOA for liver tumors induced in rodents by PPAR"

agonists is unlikely to be operative in humans (IARC, 1995). The participants of a workshop held

in 1998 under the auspices of the International Life Sciences Institute Health and Environmental

Sciences Institute concluded that although it appeared unlikely that PPAR" agonists could induce

liver tumorigenesis in humans, the possibility could not be ruled out (Cattley et al. 1998). A

recent analysis by Klaunig et al. (2003) concluded that: (1) the weight of evidence in linking

PPARα to the mode of carcinogenic action of PFOA is high for the liver; (2) the PPARα MOA is

plausible in humans since the PPARα (hPPAR") is present in humans and human livers possess

PPARα at sufficient levels to mediate the hypolipidaemic response to therapeutic fibrate drugs,

many of which are PPARα agonists; and (3) the weight of evidence, however, suggests that this

MOA is unlikely to occur in humans based on quantitative differences in several of the key

factors. For instance, human livers have been found to have 10-fold less mRNA for PPARα

compared with the rodents (Palmer et al., 1998; Tugwood et al., 1998). A recent study using a 

PPAR"-humanized mouse in which the human PPAR" was expressed in the mouse liver has

demonstrated that following activation by the potent ligand Wy-14643, the PPAR"-mediated

pathways controlling lipid metabolism are independent from those controlling the cell

proliferation pathways, and suggested that structural differences between human and mouse

PPAR" may be responsible for the differential susceptibility to liver tumor development of

fibrates (Cheung et al., 2004). 
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81

There are several additional pieces of evidence on PFOA specifically which also suggest that this

MOA is not operative in non-human primates or humans. Maloney and Waxman (1999)

demonstrated that PFOA activated the mouse PPAR" to a greater extent than the human PPARα

suggesting that humans would be less responsive to PFOA than rodents. In addition, there was no

indication of hepatic cell proliferation in a 6-month cynomolgus monkey study (Butenhoff et al.,

2002). Although limited, available human data have not shown an association between PFOA

exposure and liver effects.

On the question of the relevance of this mode of action to humans, the majority of the SAP Panel

(FIFRA SAP, 2004) agreed that there are relevant data indicating that humans are less sensitive

than rodents to the hepatic effects of PPARα agonists, and concurred with the premise that when

liver tumors are observed in long term studies in rats and mice, and 1) the data are sufficient to

establish that the liver tumors are a result of a PPAR" agonist MOA and 2) other potential MOAs

have been evaluated and found not to be operative, the evidence of liver tumor formation in

rodents should not be used to characterize potential human hepatocarcinogenic hazard.

3.9.3.3 Leydig Cell Adenomas in Rats

A series of mechanistic studies have been conducted (Cook et al., 1994; Biegel et al., 1995; Liu et

al., 1996) to investigate the MOA associated with LCT in male Sprague-Dawley rats exposed to

PFOA. Two MOAs have been proposed. One involves inhibition of testosterone biosynthesis,

leading to an imbalance of androgen/estrogen levels. This leads to an increase in leutinizing

hormone (LH) which promotes the development of LCT. The second MOA involves an increase

in serum estradiol levels via induction of hepatic aromatase activity. Estradiol stimulates the

production of growth factors such as the transforming growth factor " (TGF ") which induces

Leydig cell proliferation. 

Administration of PFOA to adult male rats by gavage for 14 days was shown to decrease

testosterone levels and increase serum estradiol levels (Cook et al., 1994). These endocrine

changes correlate 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). Other PPARα agonists (e.g., DEHP,

clofibrate) have been shown to increase serum estradiol levels in adult male rats (Eagon et al.,

1994; Rao et al., 1994). Eight out of eleven other peroxisome proliferators have been shown to

increase estradiol production using isolated LCs (Liu et al., 1996). Collectively, these data

suggest that many PPAR" agonists can increase estradiol levels. It was postulated that the

elevated estradiol levels may cause Leydig cell hyperpalsia and tumor formation by acting as a

mitogen and/or enhancing growth factor secretion; the transforming growth factor " (TGF "),

which binds to the epidermal growth factor (EGF) receptor and stimulated cell proliferation, for

instance, has been detected in Leydig cell (Teerds et al., 1990). 

Subsequent experiments have shown that PFOA increased the levels of estradiol by inducing

CYP19 (aromatase), which converts testosterone to estradiol. Peroxisome proliferators are

known to induce "-oxidation and cytochrome P-450 monooxygenases by binding to the

peroxisome proliferation activation receptor " (PPAR "; a subfamily of steroid hormone

receptors). It is possible that PFOA induces cytochrome P450 XIX ( aromatase) by binding to

and activating the PPAR ".

 

PFOA has also been shown to directly inhibit testosterone production when incubated with

isolated LCs, while ex vivo studies demonstrated that this inhibition was reversible (Biegel et al.,

1995). This inhibition of testosterone biosynthesis appears to be mediated by PPARα (Gazouli et
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82

al., 2002) and may contribute to the development of LCT through disruption of the hypothalamus

pituitary thyroid axis. Testosterone, which is synthesized and secreted by the Leydig cells, is

regulated by LH; testosterone and LH form a closed-loop feedback system in the HPT axis. In

order to maintain adequate testosterone plasma levels, reduced testosterone levels is expected to

lead to increased LH levels through the negative feedback mechanism. In a mechanistic bioassay

with PFOA, serum testosterone and LH levels were not significantly altered at the levels of PFOA

that were tested (Biegel et al., 2001). It has been pointed out that increases in LH may not always

be seen in all studies of chemicals for which the proposed mode of action calls for elevated LH,

and that compensation may have occurred to restore homeostasis and inappropriate timing of

sampling are some of the explanations for failing to detect changes in LH levels (Clegg et al.,

1997).

Thus, the above data demonstrate that the induction of LCTs by PFOA may be attributed to a

hormonal mechanism whereby PFOA either inhibits testosterone biosynthesis and/or increases

serum estradiol levels via induction of hepatic aromatase activity. A critical review of these

mode of action data by Klaunig et al. (2003) led to the conclusion that the evidence is inadequate

at this time to support a linkage between PPAR" agonism and induction of LCT. A similar

conclusion was made by the FIFRA SAP (FIFRA SAP, 2004). The Panel recognized that only

limited evidence support a link between Leydig cell tumors induction and a PPAR" agonist mode

of action, the SAP concurred with the conclusion that chemicals in this subclass that induce

Leydig cell tumors may pose a carcinogenic hazard for humans. 

Although the LCT induced in the rat by PFOA are assumed to be relevant to humans, they

probably do not represent a significant cancer hazard for humans. Testicular tumors comprise

about 1% of all human male neoplasms. LCT are rare as they account for only 1-3% of all

testicular tumors in man (cited in: Prentice and Meikle, 1995). As there is some evidence that

PFOA induced LCT in rats either by inhibiting testosterone biosynthesis and/or by inducing

aromatase and thereby increasing estradiol levels, both the rat and human hypothalamic-pituitarytesticular axis may respond to inhibition of testosterone with a subsequent increase in LH. Hence,

if PFOA inhibited testosterone biosynthesis in humans an increase in LH is expected to occur. 

However, a 6-month cynomolgus monkey study with PFOA did not demonstrate any compoundrelated effects on testosterone or estradiol (Butenhoff et al., 2002). These data suggest that PFOA

is unlikely to induce LCT in humans because humans are quantitatively less sensitive than rats to

LH stimulation (Cook et al., 1999). The number of LH receptors per Leydig cell is approximately

1,500 in man and 20,000 in the rat. Furthermore, the rat appears to be unique in possessing the

Leydig cell luteinising hormone releasing hormone (LHRH) receptors which have the same

effects on Leydig cells and are not present in man, monkey or mouse (cited in: Prentice and

Meikle, 1995). The more compelling evidence supporting that testosterone inhibition or estradiol

induction is unlikely to pose a significant cancer hazard/risk to humans may be the human disease

state FMPP where men have a mutated LH receptor that is activated throughout their life and

LCT are not seen (Cook et al., 1999). 

3.9.3.4 Pancreatic Acinar Cell Tumors in Rats

The mechanism by which PFOA induced pancreatic acinar cell tumors is unknown. A number of

other peroxisome proliferators also produce pancreatic acinar cell tumors in rats. The

development of pancreatic acinar cell hypertrophy, hyperplasia, and adenomas in the rat have

been shown to be modified by several factors such as steroids (testosterone and estradiol), growth

factors such as cholecystokinin (CCK), growth factor receptor over-expression (CCKA receptor)

and diet (fat) (Longnecker, 1983; Longnecker, 1987; Longnecker and Sumi, 1990). These
SAB Review Draft; Do Not Cite or Quote

83

potential mechanisms have been investigated in a series of in vitro and in vivo (subacute,

subchronic, and oncogenicity) studies using PFOA (Obourn et al., 1997a; Obourn et al., 1997b). 

The available data suggest that PFOA appeared to induce PACTs by an indirect trypsin-inhibition

mechanism where reduced bile flow and/or changes in bile composition produced an increase in

CCK levels secondary to hepatic cholestasis; a sustained increase in CCK levels is responsible for

the development of PACTs in rats (Cook et al., 1994; Obourn et al., 1997). CCK is a growth

factor that has been shown to stimulate normal, adaptive, and neoplastic growth of pancreatic

acinar cells in rats (Longnecker, 1987). Accordingly, the PACT induced in rats by PFOA (and

other PPAR" agonists) may be secondary to the liver effects of PPAR" agonism. However, the

mechanism(s) by which PFOA may induce PACT is based primarily on data with WY14643, a

more potent pancreatic acinar cells carcinogen. 

The available MOA data of PFOA and other PPAR" agonists which induced PACT in rats was

recently reviewed by Klaunig et al. (2003), and it was concluded that the evidence is inadequate

at this time to support a linkage between PPAR" agonism and induction of PACT. A similar

conclusion was made by the FIFRA SAP (FIFRA SAP, 2004). The SAP concurred that chemicals

in this subclass that induce pancreatic acinar cell tumors may pose a carcinogenic hazard for

humans.

While there are some evidence that a sustained increase in CCK levels is responsible for the

development of PACTs in rats by PFOA, expressions of PPARα and CCKA receptors in humans

are much lower than rodents. In addition, humans regulate pancreas exocrine secretion via a

neuronal pathway rather than direct binding of CCK to acinar CCKA receptors as in rodents.

Consistent with this conclusion, a 6-month cynomolgus monkey study with PFOA did not

demonstrate any compound-related effects on CCK levels or clinical pathology evidence of

cholestasis (Butenhoff et al., 2002). Study of CCK levels in employees also did not report

increases in workers (Olsen, et al., 1998b; Olsen et al., 2000). Furthermore, while the pancreatic

tumors in the rat are typically derived from acinar cells, the majority of human pancreatic

neoplasms are of the ductal type (Cotran et al.,1989). Therefore, the PACT induced in the rat by

PFOA probably do not represent a significant cancer hazard for humans. 

 

3.9.3.5 Cancer Descriptor

Carcinogenicity studies in Sprague-Dawley (CD) rats show that PFOA induces a “tumor triad”

similar to a number of other PPARα agonists. This “tumor triad” includes liver tumors, Leydig

cell tumors, and pancreatic acinar cell tumors. The evidence for mammary fibroadenomas in the

female rats is equivocal since the incidences were comparable to some historical background

incidences. In humans, occupational studies of still relatively young cohorts have not indicated

statistically significant increases in these types of cancer. Not only were the numbers of cancer

deaths very small in some of the cancer categories, but PFOA exposures were also not adequately 

characterized.

As summarized in section 3.9.3.1, there is sufficient evidence to conclude that PFOA is a PPAR"-

agonist and that PFOA induces liver toxicity and adenomas via a PPAR" agonist MOA in rats. 

PFOA activates the PPAR" and the requisite dose-response and/or temporal associations of the

key events for the PPAR" mode of action with the liver adenomas have been characterized. A

recent ILSI workgroup as well as the FIFRA SAP have concluded that this MOA is unlikely to

occur in humans based on quantitative differences in several of the key factors (Klaunig et al.,

2003; OPPTS, 2003; FIFRA SAP, 2004). 
SAB Review Draft; Do Not Cite or Quote

84

The modes of carcinogenic action of PFOA-induced LCTs and PACT have not been fully

elucidated. The induction of LCT by PFOA may involve a hormonal mechanism whereby PFOA

either inhibits testosterone biosynthesis and/or increases serum estradiol levels via induction of

hepatic aromatase activity. The induction of PACT are related to an increase in serum level of the

growth factor, CCK, that appears to be secondary to changes in the liver. As the modes of

carcinogenic action of PFOA-induced LCT and PACT have not been clearly linked to PPAR"

agonism, it is assumed that these tumors induced in rats are relevant to humans. However, the

LCT and PACT induced in the rat by PFOA probably do not represent a significant cancer hazard

for humans because of quantitative differences in the expressions of LH and CCKA receptors and

of other toxicodynamic differences between the rat and the human. For instance, the number of

LH receptors per Leydig cell is approximately 1,500 in man and 20,000 in the rat. Furthermore,

the rat appears to be unique in possessing the Leydig cell luteinising hormone releasing hormone

(LHRH) receptors which have the same effects on Leydig cells and are not present in man,

monkey or mouse (cited in: Prentice and Meikle, 1995). While there is some evidence that a

sustained increase in CCK levels is responsible for the development of PACTs in rats by PFOA,

expressions of PPARα and CCKA receptors in humans are much lower than rodents. In addition,

humans regulate pancreas exocrine secretion via a neuronal pathway rather than direct binding of

CCK to acinar CCKA receptors as in rodents; the pancreatic tumors in the rat are typically derived

from acinar cells whereas the majority of human pancreatic neoplasms are of the ductal type

(Cotran et al.,1989).

Overall, based on no adequate human studies and uncertain human relevance of the tumor triad

(liver, Leydig cell and pancreatic acinar cell tumors) from the rat studies, PFOA may be best

described as “suggestive evidence of carcinogenicity, but not sufficient to assess human

carcinogenic potential”under the draft 1999 Guidelines for Carcinogen Risk Assessment (U.S.

EPA, 1999).

3.9.4 Toxicity in Adult Repeat-Dose Animal Studies

3.9.4.1 Non-Human Primates

A summary of the toxicity associated with exposure to APFO in adult non-human primates is

presented in Table 21. In both the Rhesus and cynomolgus monkey studies, there was a very

steep dose-response curve for mortality. The cause of the mortality is unknown. The Rhesus

monkeys also exhibited several clinical indicative of gastrointestinal distress; the mode of action

for these effects is unknown, but the gastrointestinal effects may be due to the potent surface

activity properties of APFO. The gastrointestinal effects were not noted in a 6 month study of

cynomolgus monkeys. The lack of gastrointestinal effects may be due to a species difference, or

due to the fact that the monkeys were dosed via a capsule versus gavage in the Rhesus monkey

study. In addition, liver weights were significantly increased in the cynomolgus monkey, but

were unaffected in the Rhesus monkey. The increased liver weight does not appear to be due to

peroxisome proliferation as there was only a minimal increase in palmitoyl CoA oxidase in the

high dose group. However, there was evidence of mitochondrial proliferation which also

suggests that a mode of action other than PPAR"-agonism is operative in this species. It should

also be noted that a number of hormonal and cell proliferation assays were included in the

cynomolgus monkey study to determine whether there was any indication of a response of

hepatic, Leydig cells and pancreatic acinar cells to PFOA exposure which have been observed in

the rat. There was no indication of cell proliferation in the liver, testes or pancreas, nor were

hormonal measures affected. There is significant uncertainty associated with the derivation of a

LOAEL and NOAEL for both of these studies due to the small sample size, the variability in
SAB Review Draft; Do Not Cite or Quote

85

response of the individual monkeys, and the lack of linear proportionality between administered

dose and serum PFOA levels.

Table 21

Summary of Toxicity in Nonhuman Primates

Monkey Exposure

Period Dose Response

LOAEL/

NOAEL

(mg/kg-day) Reference

Rhesus

2/sex/dose

13 weeks gavage

0, 3, 10, 30,

100 mg/kg-day

> 3 - clinical signs

(including G.I)

> 30 - mortality, 9

body weight, clinical

signs

LOAEL = 3

no NOAEL

Goldenthal,

1978b

Cynomolgus

4-6

males/dose

6 months oral capsule 

0, 3, 10, 20/30

mg/kg-day

3 - moribund

(sacrificed), clinical

signs

> 3 - 8absolute and

relative liver weight

20/30 - moribund

(sacrificed), clinical

signs, treatment

stopped, 9body

weight, 

LOAEL = 3

NOAEL =

ND

Thomford,

2001b;

Butenhoff et

al., 2002

3.9.4.2 Adult Male Rats

A summary of the systemic toxicity associated with APFO exposure in adult male rats is

presented in Table 22. The major target organ is the liver. As described in section 3.9.2.1, the

liver toxicity is due to a PPAR"-agonist mode of action which is unlikely to occur in humans. 

Aside from the liver toxicity, APFO exposure is not associated with many other effects. Body

weight was reduced at doses of 6.5 mg/kg-day in a 90-day study (Palazzolo, 1993); at > 3 mg/kgday in the F0 males and > 1 mg/kg-day in the F1 males in the two-generation reproductive

toxicity study (York, 2002; Butenhoff et al., 2004); and at 14.2 mg/kg-day in the 2-year study

(Sibinski, 1987). Adrenal hypertrophy and vacuolation was noted in the F0 adult males at doses

> 10 mg/kg-day and in F1 adult males at 30 mg/kg-day in the two-generation reproductive

toxicity study (York, 2002; Butenhoff et al., 2004). There is some uncertainty as to the biological

significance of this finding since this effect was not observed in a 90-day study (Goldenthal,

1978a) or a 2-year study (Sibinski, 1987). Exposure to APFO was associated with a reduction in

the number of erythrocytes and related parameters following exposure to 14.2 mg/kg-day APFO

in the 2-year study (Sibinski, 1987). The testicular masses and testicular vascular mineralization

observed in the 2-year study are probably associated with the etiology of the Leydig cell

adenomas that were discussed in section 3.9.4.3. The mode of action is unknown for any of these

effects.
SAB Review Draft; Do Not Cite or Quote

86

Table 22

Summary of Systemic Toxicity in Adult Male Rats

Rat 

Strain

Exposure

Period Dose

Effects

(mg/kg-day)

LOAEL/

NOAEL

(mg/kg-day) Reference

CD 13 weeks diet - 0, 10,

30, 100, 300,

1000 ppm

0.056, 1.72,

5.64, 17.9, 

63.5 mg/kg-day

> 1.72 - 8absolute liver

 weight

> 5.64 - 8hepatocellular

hypertrophy

> 17.9 - 8relative liver

weight

> 63.5 - 9body weight

LOAEL 

= 1.72

NOAEL

 = 0.56

Goldenthal,

1978a

SpragueDawley13 weeks diet - 0, 1, 10,

30, 100 ppm

0.06, 0.64, 1.94, 

6.50 mg/kg-day

> 0.64 - 8absolute and

relative liver weight,

hepatocellular 

hypertrophy,

8palmitoyl CoA oxidase

6.5 - 9body weight

LOAEL 

= 0.64

NOAEL 

= 0.06

Palazzolo,

1993

SpragueDawleyF0 animals

-15 weeks

F1 animals

-19-20 weeks

gavage - 

0, 1, 3, 10,

30 mg/kg-day

> 1 - 8absolute and

relative liver

> 3 - 9body weight

> 10 - adrenal

hypertrophy and

vacuolation

30 - clinical signs,

1/30 moribund

>1 - 8absolute and

relative liver weight,

9body weight

>3 - hepatocellular

hypertrophy

30 - adrenal

hypertrophy and

vacuolation

LOAEL 

=1

no NOAEL

LOAEL 

=1

no NOAEL 

York, 2002;

Butenhoff

et al., 2004

SpragueDawley2 years diet - 0, 30,

300 ppm

1.3, 14.2

mg/kg-day

14.2 - 8relative liver

weight, hepatocellular

hypertrophy, cystoid

degeneration, ALT;

9 body weight gain;

9erythrocytes,

hematocrit,

hemoglobin

concentration;

testicular vascular

mineralization,

testicular masses

LOAEL 

= 14.2

NOAEL 

= 1.3

Sibinski, 

1987

SpragueDawley2 years diet -

0, 300 ppm

14.2 mg/kg-day

14.2 - 8 relative

liver weight

LOAEL 

= 14.2

no NOAEL

Cook et al.,

1994;

Biegel et al.,

2001
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87

3.9.4.3 Adult Female Rats

A summary of the systemic toxicity associated with APFO exposure in the adult female rat is

shown in Table 23. Due to the gender difference in elimination of PFOA in rats, the response of

the female rat to APFO exposure is generally far less than that of the male rat. The effects on the

liver were only observed at relatively high doses (Goldenthal, 1978a). The effects on the adrenal

that were observed in the male rat were not observed in the female rat in the two-generation

reproductive toxicity study (York, 2002; Butenhoff et al., 2004). Body weight was reduced at

doses of 30 mg/kg-day in the F1 females in the two-generation reproductive toxicity study (York,

2002; Butenhoff et al., 2004), and at 16.1 mg/kg-day in the 2-year study (Sibinski, 1987). 

Exposure to APFO was associated with a reduction in the number of erythrocytes and related

parameters following exposure to 16.1 mg/kg-day APFO in the 2-year study (Sibinski, 1987). 

The mode of action is unknown for any of these effects.

Table 23

Summary of Systemic Toxicity in Adult Female Rats

Rat 

Strain

Exposure

Period Dose

Effects

(mg/kg-day)

LOAEL/

NOAEL

(mg/kg-day) Reference

CD 13 weeks diet

0, 10, 30, 100,

300, 1000 ppm

0.74, 2.3, 7.7,

22.4, 76.5

mg/kg-day

76.5 - 8absolute and

relative liver weight

LOAEL

 = 76.5

NOAEL 

= 22.4

Goldenthal,

1978a

SpragueDawleyF0 animals

-18 weeks

F1 animals

-18 weeks

gavage

0, 1, 3, 10, 30

mg/kg-day

No effects

30 - 9 body weight

no LOAEL 

NOAEL 

= 30

LOAEL 

= 30

NOAEL 

= 10

York, 2002;

Butenhoff

et al., 2004

SpragueDawley2 years diet 

0, 30, 300 ppm

1.6, 16.1

mg/kg-day

16.1 - 9body weight

gain, 9erythrocytes,

hematocrit, hemoglobin

concentration

LOAEL 

= 16.1

NOAEL

= 1.6

Sibinski,

1987

3.9.4.4 Adult Mice

In mice, a 28 day study demonstrated that the liver is also a target organ. As described above

(section 3.9.3), the liver toxicity is due to a PPAR"-agonist mode of action which is unlikely to

occur in humans. In addition, several studies have shown that PFOA affects the immune system

in mice. Feeding C57Bl/6 mice a diet containing 0.02% PFOA resulted in adverse effects to both

the thymus and spleen (Yang et al. 2000, 2001). In addition, this feeding regimen resulted in
SAB Review Draft; Do Not Cite or Quote

88

suppression of the specific humoral immune response to horse red blood cells, and suppression of

splenic lymphocyte proliferation in response to LPS and ConA (Yang et al., 2002b). The

suppressed mice recovered their ability to generate a humoral immune response when they were

fed a diet devoid of PFOA. Studies using transgenic mice showed that the PPAR" was involved

in causing the adverse effects to the immune system (Yang et al., 2002a). Further research is

needed to determine the exact role of PPAR" and immunotoxicity in mice.

3.9.5 Developmental and Reproductive Toxicity in Animal Studies

The developmental and reproductive effects associated with oral exposure to APFO in animal

studies are summarized in Table 24. In New Zealand White rabbits, there was evidence of

prenatal developmental toxicity following exposure to 50 mg/kg-day APFO on gestation days 6-

18 (Gortner, 1982). In contrast, there was no evidence of developmental toxicity following

exposure to doses as high as 150 mg/kg-day APFO during gestation days 6-15 in Sprague-Dawley

rats (Gortner, 1981; Staples, 1984). However, there was evidence of developmental/reproductive

toxicity during the postnatal period in a two generation reproductive toxicity study in SpragueDawley rats (York, 2002; Butenhoff et al., 2004). In this study, During lactation, there was a

reduction in F1 mean body weight on a litter basis during lactation (sexes combined) in the 30

mg/kg-day group. F1 males in the 10 and 30 mg/kg-day groups exhibited a significant reduction

in body weight gain during days 8-50 postweaning, and body weights were significantly reduced

in the 10 mg/kg-day group beginning on postweaning day 36, and in the 30 mg/kg-day group

beginning on postweaning day 8. F1 females in the 30 mg/kg-day group exhibited a significant

reduction in body weight gain on days 1-15 postweaning, and in body weights beginning on day 8

postweaning. There was a significant increase in mortality mainly during the first few days after

weaning, and a significant delay in the timing of sexual maturation for F1 males and females in

the 30 mg/kg-day group. For F1 males, the LOAEL for developmental/reproductive toxicity was

10 mg/kg-day, and the NOAEL was 3 mg/kg-day. For F1 females, the LOAEL for

developmental/reproductive toxicity was considered to be 30 mg/kg-day, and the NOAEL was 10

mg/kg-day. No effects were observed in the F2 pups. However, it should be noted that the F2

pups were sacrificed at weaning, and thus it was not possible to ascertain the potential postweaning effects that were noted in the F1 generation. 
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89

Table 24

Summary of Developmental and Reproductive Toxicity Effects

Species

Dose

(mg/kg-day)

Exposure 

Duration

Time

Assessed Effects

LOAEL/

NOAEL Reference

New Zealand

White rabbit

0, 1.5, 

5, 50 

GD 6-18 GD 29 813th rib LOAEL

= 50

NOAEL 

= 5

Gortner, 1982

SpragueDawley

rat

0, 0.05, 

1.5, 5, 150 

GD 6-15 GD 20 none no LOAEL

NOAEL 

= 150

Gortner, 1981

SpragueDawley

rat

0, 100 GD 6-15 GD 20 none no LOAEL

NOAEL 

= 100

Staples, 1984

SpragueDawley

rat

0, 100

GD 6-15 PND 35

none no LOAEL

NOAEL 

= 100

Staples, 1984

SpragueDawley

rat 

0, 1, 3,

10, 30

F1 - prior to

conception

through 

sexual

maturation

F2 - prior to

conception

until 

weaning

Variety

Variety

> 10 - 

9postweaning

body weight 

in males

30 - 

9preweaning

litter body

weight ,

8postweaning 

mortality,

delayed sexual

maturation,

9postweaning

body weight 

in females

none

LOAEL 

= 10

NOAEL 

= 3

no LOAEL

NOAEL 

= 30

York, 2002;

Butenhoff et

al., 2004
SAB Review Draft; Do Not Cite or Quote

90

4.0 Biomonitoring Data

PFOA has been measured in human serum of workers occupationally exposed to APFO and in the

general U.S. population. In general, serum PFOA concentrations in the general population are

much lower than in workers exposed to APFO (see Tables 25-27 below). However, it should be

noted that the highest levels reported to date in the general population are similar to some of the

lowest levels in workers exposed to PFOA occupationally. The environmental concentrations of

APFO and the pathways of exposure to the general population are not known.

4.1 Occupational Exposures

3M and DuPont have been the primary producers and users of perfluorinated compounds in the

U.S. Both companies offer voluntary medical surveillance programs to workers at plants that

produce or use perfluorinated compounds. 3M discontinued manufacturing PFOA between 2000-

2002. PFOA serum levels seem to be decreasing slightly at both the 3M and DuPont facilities;

however, the cross-sectional nature of the data cannot provide trends in serum concentrations. 

PFOA blood serum data for volunteers in the medical surveillance programs at these plants that

have been submitted to EPA are presented below in μg/mL (ppm).

4.1.1 3M Occupational Data

3M has been offering voluntary medical surveillance to workers at plants that produce or use

perfluorinated compounds since 1976. Serum PFOA levels in 3M workers have been measured

since 1993 (Olsen et al., 2003a; 2003b; 2003c; 2001a; 2001b; 2001c; 2000; 1999; 1998c). Prior

to this time, analytical capabilities precluded the accurate measurement of any specific

fluorochemical analyte and only total organic fluorine was measured. PFOA analysis differed

slightly in each surveillance program year and different laboratories were used to assay PFOA in

each year. The samples were analyzed for PFOA using high performance liquid chromatography

mass spectrometry (HPLC/MS), but the extraction methods differed slightly each year (Olsen, et

al., 2001f; Olsen, et al., 2003a; 2003b). 

Serum PFOA concentrations for workers participating in 3M’s biomonitoring program have been

reported for 3 plants: Cottage Grove, Minnesota; Decatur, Alabama;and Antwerp, Belgium. 

Only the U.S. data are presented here. Surveillance years included 1993, 1995, 1997, 1998, 2000,

2002, and 2003, although not all of the plants offered surveillance in all of these years. 

Participation in all of the program years was voluntary. The eligible voluntary participation rates

ranged from approximately 70% in 1993 to 50% in 1997 (Olsen, et al., 2001f). The number of

employees volunteering for participation increased greatly in 2000 and then declined by about

75% in 2002 (Olsen, et al., 2003a; 2003b) (see Table 24 below). 3M suggests that the increase in

2000 was probably due to increased employee awareness of the persistence and prevalence of

perfluorinated chemicals in human tissues. Meanwhile, the decline in 2002 was probably due to

3M’s announcement to phase out PFOA production at their plants. In 1998, a random sample of

Decatur employees was surveyed to determine whether the voluntary nature of the program was

biasing the results of the surveillance (Olsen et al., 1999). The data presented in Table 24

indicate that there is little difference in male serum concentrations between 1997 (1.40 μg/mL,

when participation was voluntary) and 1998 (1.735 μg/mL, when a random sample was taken). 

Gender-based data for 3M’s biomonitoring program are limited due to the small number of

women participating in the program. Mean PFOA serum concentrations for female employees at
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91

all 3 of the plants were lower than those of male employees, even in similar jobs. Where data are

available by gender, they are provided in Table 25.

Table 25

Summary of 3M U.S. Occupational PFOA Biomonitoring 

(serum concentrations, μg/mL)*

Plant

Arithmetic

Mean

Range

Geometric

Mean 

95% CI of

Geometric

Mean

Reference

Cottage Grove

2002 (n=38)

2000 (n = 148)

 Male (n = 131)

 Female (n = 17)

1997 (n =74)

1995 (n = 80)

1993 (n = 111)

4.3

NR

4.51

0.85

6.4

6.8

5.0

0.07 - 32.6

NR

.007-92.03

.04 - 4.73

0.1 - 81.3

0 - 114.1

0 - 80.0

1.7

NR

0.85

0.42

NR

NR

NR

1.02-2.72

NR

0.64- 1.22

0.23 - 0.79

NR

NR

NR

Olsen, et al., 2003a

Olsen, et al., 2003f

Olsen, et al., 2000

Olsen, et al., 2000

Olsen, et al., 2000

Decatur

2002 (n =54)

2000 (n = 263)

 Male (n = 215)

 Female ( n = 48)

1998 (n = 126)

 Male (n =102) 

 Female (n=24)

1997 (n = 84)

 (males only)

1995 (n = 90)

 (males only)

1.497

1.78

1.90

1.23

1.54

1.735

0.691

1.40

1.72

0.025-4.81

0.04-12.70

0.04-12.70

NR

0.02 - 6.76

0.02 - 6.76

NR

NR

NR

0.713

1.13

NR

NR

0.90

1.142

0.326

NR

NR

0.483-1.055

0.99 - 1.30

NR

NR

0.72 -1.12

NR

NR

NR

NR

Olsen, et al., 2003b

Olsen, et al., 2001a

Olsen et al., 2003e

Olsen et al., 2003e

Olsen, et al., 1999

Olsen, et al., 1998c

Olsen, et al., 1998c

Building 236

(fluorochemical

research) 2000 

(n = 45)

0.106 0.008-

0.668

0.053 0.037 -0.076 Olsen, et al., 2001c

*μg/mL = ppm

NR- not reported

Of the plants listed above, PFOA exposures are highest at the Cottage Grove plant. 

Correspondingly, PFOA serum levels are highest at this plant (geometric mean = 1.7 μg/mL,

range, 0.07 - 32.6 μg/mL). Mean serum PFOA levels increased slightly at both the Cottage

Grove and Decatur plants from 1993 to 2000; however, the latest serum measurements (2002)

show a slight decline in mean PFOA concentrations. However, it should be noted that the

number of participants in the 2002 sampling period was comparatively small. In addition, all of

the data are cross-sectional and therefore cannot provide any temporal information. 
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92

4.1.2 DuPont Occupational Data

Dupont has been measuring PFOA in workers’ blood serum since 1981 (Dupont, 2001a, 2001b). 

Prior to 1981, total blood fluoride levels were reported. All of Dupont’s operations in the U.S.

that use PFOA with significant exposure potential are concentrated at the Washington Works

plant in Washington, West Virginia. PFOA is used as a fluoropolymer reaction aid at this plant. 

The data presented in Table 26 reflect serum concentrations of volunteer workers in the plant who

had potential PFOA exposure. With the exception of the 2000 sampling period, the same workers

were common to all of the other sampling periods, with additional workers participating in 1995. 

The data available to EPA at this time are limited to what is presented here.

Table 26 

Summary of Dupont Occupational PFOA Biomonitoring 

(serum concentrations, μg/mL)*

Year

Arithmetic

Mean Range Reference

Washington Works

2000 (n= 72)

1995 (n = 80)

1989-90 (n =22 )

1985 (n = 22)

1984 (n = 19)

1.53

1.56

1.96

2.34

3.21

0.02 - 9.0

0.12 - 4.5

0.06 - 11

0.06 - 18

0.07 - 24

Dupont, 2001a

Dupont, 2001b

Dupont, 2001b

Dupont, 2001b

Dupont, 2001b

*μg/mL = ppm

4.2 General Population Exposures

Data on PFOA levels in the general population include both pooled and individual serum

samples. Mean serum PFOA levels are lower in the general population than in workers exposed

to PFOA (described above). Individual serum PFOA levels in three separate US cohorts are very

similar. The geometric mean for all three cohorts, comprised of 3 separate age groups, is

approximately 0.004 μg/mL (4 ppb). All of the data are fairly recent but are cross-sectional;

therefore, temporal trends cannot be established. The available data are summarized in Table 27.
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93

Table 27

Summary of General Population PFOA Biomonitoring 

(serum concentration, μg/mLa)

Sample

Arithmetic

Mean Range

Geometric

Mean 95% CI

Reference

Pooled Samples

Commerical sources of

blood, 1999 (n = 35 lots) 0.003 .001 -

.013

NA NA

3M

Company,

1999a

Blood banks, 1998 (n =

18 lots, 340-680 donors) 0.017b 0.012 -

0.022

NA NA

3M

Company,

1999b

Individual Samples

American Red Cross

blood banks, 2000 (n =

645)

0.0056 0.0019 -

0.0523

0.0046 0.0043 -

0.0048

Olsen et al.,

2002a,

2003d

Elderly (65 - 96 years),

2000 (n = 238) NR 0.0014 -

0.0167

0.0042 0.0039 -

0.0045

Olsen et al.,

2002b,

2004a

Children (2 - 12 years),

1995 (n = 598) 0.0056 0.0019 -

0.0561

0.0049 0.0047 -

0.0051

Olsen et al.,

2002c,

2004b

a

 μg/mL = ppm 

b PFOA detected in about 1/3 of the pooled samples but quantifiable in only 2.

NR - not reported

Pooled blood samples from U.S. blood banks indicate mean PFOA levels of 0.003 to 0.017

μg/mL (3 - 17 ppb) ( 3M Company, Feb. 5, 1999; 3M Company, May 26, 1999). The highest

pooled sample reported was 0.022 μg/mL (22 ppb). Pooled samples were collected in 1998 and

1999. However, it cannot be assumed that these levels can be generalized to the U.S. population

for several reasons: 1) blood donors are a unique group that does not necessarily reflect the U.S.

population as a whole, 2) many of the blood banks originally contacted for possible inclusion in

the study declined to participate, 3) only a small number of samples have actually been analyzed

for PFOA, and 4) no other data such as age, sex, or other demographic information are available

on the donors.

Individual blood samples from 3 different age populations were recently analyzed for PFOA and

other fluorochemicals using high-pressure liquid chromatography/electrospray tandem mass

spectrometry (HPLC/ESMS) (Olsen et al., 2002a, 2002b, 2002c). The studies’ participants

included adult blood donors, an elderly population participating in a prospective study in Seattle,
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94

WA, and children from 23 states participating in a clinical trial. Overall, the PFOA geometric

means were similar across all 3 populations (0.0046 μg/mL, 0.0042 μg/mL, and 0.0049 μg/mL,

respectively). The geometric means and 95% tolerance limits (the exposure below which 95% of

the population is expected to be found) and their upper bounds were comparable across all 3

studies. However, the upper ranges for the children and adults were much higher than for the

elderly population. It is not clear whether this is the result of geographic differences in PFOA

levels or some other factor. It should be noted that PFOS and PFOA were highly correlated in all

three studies (r = .63, r = .70, and r = .75) and that PFOA did not meet the criteria for a log

normal distribution based on the Shapiro-Wilk test in any of the studies. The authors suggest that

it may be due to the greater proportion of subjects with values less than the lower limit of

quantitation (LLOQ); however, only 12 of the 1481 total samples were below the LLOQ. In

those instances where a sample was measured below the LLOQ, the midpoint between zero and

the LLOQ was used for calculation of the geometric mean. The details of each study are

provided below.

Blood samples from 645 U.S. adult blood donors (332 males, 313 females), ages 20-69, were

obtained from six American Red Cross blood banks located in: Los Angeles, CA;

Minneapolis/St. Paul, MN; Charlotte, NC; Boston, MA; Portland, OR, and Hagerstown, MD

(Olsen et al., 2002a). Each blood bank was requested to provide approximately 10 samples per

10-year age intervals (20-29, 30-39, etc.) for each sex. The only demographic factors known for

each donor were age, gender, and location.

The geometric mean serum PFOA level was 0.0046 μg/mL. The range was <lower limit of

quantitation (0.0019 μg/mL) to 0.0523 μg/mL. Only 2 samples were less than the LLOQ. Males

had significantly higher (p < .05) geometric mean PFOA levels than females (0.0049 μg/mL vs.

0.0042 μg/mL). Age was not an important predictor of adult serum fluorochemical

concentrations. When stratified by geographic location, the highest geometric mean for PFOA

was in the samples from Charlotte, NC (0.0063 μg/mL, range: 0.0021 – 0.029 μg/mL) and the

lowest from Portland (0.0036 μg/mL, range: 0.0021 – 0.0167 μg/mL). The highest individual

value was reported in Hagerstown (0.0523 μg/mL). 

Serum PFOA levels were reported for 238 (118 males and 120 females) elderly volunteers in

Seattle participating in a study designed to examine cognitive function in adults aged 65-96

(Olsen et al., 2002b, 2004a). Age, gender and number of years’ residence in Seattle were the only

data available on the participants. Most of the participants were under the age of 85 and had lived

in the Seattle area for over 50 years.

The geometric mean of PFOA for all samples was 0.0042 μg/mL. The range was 0.0014 –

0.0167 μg/mL. Only 5 samples were less than the LLOQ of 0.0014 μg/mL. There was no

significant (p < .05) difference in geometric means for males and females. In simple linear

regression analyses, age was negatively (p < .05) associated with PFOA in elderly men and

women. PFOS and PFOA were highly correlated (r = .75) in this study.

A sample of 598 children, ages 2-12 years old, participating in a study of group A streptococcal

infections, was analyzed for serum PFOA levels (Olsen et al., 2002c, 2004b). The samples were

collected in 1994-1995 from children residing in 23 states and the District of Columbia. PFOA

did not meet the criteria for a log normal distribution based on the Shapiro-Wilk test. The authors

suggest that it may be due to the greater proportion of subjects with values < LLOQ for PFOA;

however, only 5 samples were less than the LLOQ of 0.0019 μg/mL. The geometric mean of

PFOA for all of the participants was 0.0049 μg/mL, and the range was 0.0019 to 0.0561 μg/mL. 
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95

Male children had significantly (p<.01) higher geometric mean serum PFOA levels than females: 

0.0052 μg/mL and 0.0047 μg/mL, respectively. In simple linear regression analyses, age was

significantly (p < .05) negatively associated with PFOA in both males and females. When

stratified by age, the geometric mean of PFOA was highest at age 4 (0.0057 μg/mL) and lowest at

age 12 (0.0035 μg/mL). Although the data were not reported, a graphical presentation of log

PFOA levels for each state by gender looked similar across the states. However, it is difficult to

interpret these data without analyzing them and the sample sizes were limited for each

gender/location subgroup. PFOS and PFOA were highly correlated (r = .70) in this study. PFOA

and PFHS (perfluorohexanesulfonate) were also correlated, although not as strongly (r = .48). 

The above 3 studies indicate similar geometric means and ranges of PFOA among sampled

adults, children, and an elderly population. However, an unexpected finding was the level of

PFHS and M570 (N-methyl perfluorooctanesulfonamidoacetate) in children (Olsen, 2002c,

2004b). These serum levels were much higher in the sampled children than in the sampled adults

or elderly. It is not clear why this occurred, but it is probably due to a different exposure pattern

in children.

In another study, the PFOA concentration was analyzed in human sera and liver samples using

HPLC/ESMS (Olsen et al., 2001d). Thirty-one donor samples were obtained from 16 males and

15 females. The average age of the male donors was 50 years (SD 15.6, range 5-69) and the

average age of the female donors was 45 years (SD 18.5, range 13-74). The causes of death were

intracranial hemorrhage (n = 16 or 52%), motor vehicle accident (n = 7 or 23%), head trauma (n =

4 or 13%), brain tumor (n = 2 or 6%), drug overdose (n = 1 or 3%) and respiratory arrest (n = 1 or

3%). Both serum and liver tissue were obtained from 23 donors; 7 donors contributed liver tissue

only and 1 donor contributed serum only. Resulting serum values for PFOA ranged from < LOQ 

(<0.0030) - 0.0070 μg/mL. Assuming the midpoint value between zero and LOQ serum value for

samples <LOQ, the mean serum PFOA level was calculated as 0.0031 μg/mL with a geometric

mean of 0.0025 μg/mL. No liver to serum ratios were provided because more than 90% of the

individual liver samples were <LOQ.

5.0 Risk Assessment

A margin of exposure (MOE) approach can be used to describe the potential for human health

effects associated with exposure to a chemical. The MOE is calculated as the ratio of the

NOAEL or LOAEL for a specific endpoint to the estimated human exposure level. The MOE

does not provide an estimate of population risk, but simply describes the relative “distance”

between the exposure level and the NOAEL or LOAEL. In this risk assessment there is no

information on the sources or pathways of human exposure. However, serum levels of PFOA,

which are a measure of cumulative exposure, were available from human biomonitoring studies. 

In addition, serum levels of PFOA were available for many of the animal toxicology studies or

there was sufficient pharmacokinetic information to estimate serum levels. Thus, in this

assessment internal doses from animal and human studies were compared; this is somewhat

analogous to a MOE approach which uses external exposure estimates.

5.1 Selection of Endpoints

The epidemiology and animal toxicology studies are summarized in section 3.4.9. The results of

existing epidemiology studies are not adequate for use in quantitative risk assessment, and

therefore the analysis is restricted to endpoints in the animal toxicology studies. A variety of

studies have examined the potential toxicity of PFOA in adult animals. 
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96

Studies in mice have shown an effect on the immune system. These mouse studies were not

considered for the calculation of the MOEs since information on serum levels of PFOA was not

available, nor are there pharmacokinetic data that would allow for the estimation of serum levels. 

As summarized in section 3.9.5.1, studies of non-human primates did not establish a NOAEL. 

The LOAEL for the 13-week study in Rhesus monkeys was 3 mg/kg-day based on clinical signs

of toxicity (Goldenthal, 1978b) and the LOAEL for the 6-month study of cynomolgus monkeys

was also 3 mg/kg-day based on increased liver weight and possibly mortality (Thomford, 2001b;

Butenhoff et al., 2002). For calculation of MOEs, the cynomolgus monkey study was chosen

since it included a slightly larger sample size (N = 4 to 6), was of longer duration, and included

more robust serum analyses of PFOA.

Several studies have examined the toxicity of APFO in adult male and female rats which are

summarized in sections 3.9.5.2 and 3.9.5.3, respectively. In adult male rats, the primary target

organ is the liver. However, as described in section 3.9.3, the proposed mode of action for PFOA

(PPAR"-agonism) in rodents leading to the observed liver toxicity is unlikely to occur in humans,

and therefore was not considered for calculation of the MOEs. The only other consistent effect

noted across studies was body weight reduction. Body weight was reduced at doses of 6.5 mg/kgday in a 90-day study (Palazzolo, 1993); at > 3 mg/kg-day in the F0 males and > 1 mg/kg-day in

the F1 males in the two-generation reproductive toxicity study (York, 2002; Butenhoff et al.,

2004); and at 14.2 mg/kg-day in the 2-year study (Sibinski, 1987). For adult female rats, there are

very few effects due to the rapid excretion of PFOA. The only consistent effect that was noted

among the studies was reduced body weight; this was observed at doses of 16.1 mg/kg-day in the

2-year study (Sibinski, 1987) and at 30 mg/kg-day in the adult F0 and F1 females in the twogeneration reproductive toxicity study (York, 2002; Butenhoff et al., 2004). Although, the

significance of reduced body weight for human health is not clear, body weight was used for the

calculation of the MOEs simply because it defines the lowest effect level and therefore serves as a

conservative benchmark. The lowest dose that was associated with reductions in body weight in

males was in the F1 males in the two-generation reproductive toxicity study. For the F1 males,

the LOAEL for body weight is 1 mg/kg-day, and a NOAEL was not established. The lowest dose

for females was in the 2-year study; the LOAEL for body weight was 16.1 mg/kg-day and the

NOAEL was 1.6 mg/kg-day.

Several studies have examined the developmental and reproductive toxicity associated with

exposure to APFO which are summarized in section 3.9.5.5. Developmental effects were

observed in a prenatal developmental toxicity study in New Zealand White rabbits (Gortner,

1982). However, this study was not considered for the calculation of the MOEs since information

on serum levels of PFOA in rabbits was not available, nor are there pharmacokinetic data that

would allow for the estimation of serum levels in rabbits. There were no developmental effects

noted in prenatal developmental toxicity studies of Sprague-Dawley rats following exposure to

APFO during gestation days 6-15 (Gortner, 1981; Staples, 1984). Developmental effects were

noted in the F1 male and female pups in the two-generation reproductive toxicity study. During

lactation, there was a reduction in F1 mean body weight on a litter basis (sexes combined) in the

30 mg/kg-day group. F1 males in the 10 and 30 mg/kg-day groups exhibited a significant

reduction in body weight gain during days 8-50 postweaning, and body weights were

significantly reduced in the 10 mg/kg-day group beginning on postweaning day 36, and in the 30

mg/kg-day group beginning on postweaning day 8. F1 females in the 30 mg/kg-day group

exhibited a significant reduction in body weight gain on days 1-15 postweaning, and in body

weights beginning on day 8 postweaning. There was a significant increase in mortality mainly

during the first few days after weaning, and a significant delay in the timing of sexual maturation
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97

for F1 males and females in the 30 mg/kg-day group. These endpoints were used for the

calculation of the MOEs.

5.2 Use of Serum Levels as a Measure of Internal Dose for Humans

5.2.1 General Population

A summary of the human serum levels of PFOA that were considered in the calculation of MOEs

is provided in Table 28. These data are derived from two biomonitoring studies performed on

large cohorts from various geographic areas of the US and include both children and adults. They

are described in detail in Section 4.2. The arithmetic and geometric means, as well as the range,

95% confidence intervals for the geometric means, and the cumulative 90th percentiles are

displayed in order to provide a sense of the distribution of the data. The frequency distributions

of the serum data appear to be log-normally distributed, although the criteria for the Shapiro-Wilk

test are not met. This may be due to the number of subjects with serum values less than the

LLOQ (n = 2 for <1.9 ng/ml and n = 48 for < 2.1 ng/ml). Gender specific data were available for

the geometric mean and range, but not for the arithmetic mean. These serum values are assumed

to represent steady state levels in the general US population.

Table 28 

Summary of Levels of PFOA (μg/mL) in the Serum of Human Populations

Population

Arithmetic

Mean

90th

percentile Range

Geometric

Mean

(GM)

95% CI

GM

Adults (20 - 69 years, American

Red Cross blood banks, 2000) 

 Both genders (n = 645)

 Males (n = 332)

 Females (n = 313)

0.0056 

NR

NR

0.0094

0.0101

0.0084

0.019-0.0523

<0.0019-0.029 

<0.0021-0.0523

0.0046 

0.0049 

0.0042

0.0043-0.0048

0.0046 - 0.0053

0.0039 - 0.0045

Children (2-12-years, 1995)

 Both genders (n = 598)

 Males (n = 300)

 Females (n = 298)

0.0056 

0.0059

0.0052

0.0085

0.0090

0.0080

0.0019 - 0.0561

<0.0029- 0.0561

<0.0019- 0.0186

0.0049 

0.0052

0.0047

0.0047 - 0.0051 

0.0049 - 0.0052

0.0044 - 0.0049

Unless stated otherwise, the geometric mean and 90th percentiles for the combined male and

female adults and combined male and female children are used in the calculations of the MOEs in

the following sections. 

5.2.2 Workers

Although some serum level data were available for workers, the data were not adequate to

calculate MOEs for specific occupational exposures. As described in section 4.1, 3M and DuPont

have provided serum monitoring to workers on a voluntary basis. 3M discontinued

manufacturing PFOA between 2000-2002. The last serum monitoring was offered at the 3M

plants in 2002, but there were about 75-80% fewer volunteers in 2002 than in 2000. Thus, the

sample size for 2002 is very small. In addition, since occupational exposures no longer exist for

3M workers, use of the blood monitoring data for 2000 may overestimate current serum levels. 
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98

The data available for the DuPont workers was very limited and there was no information

available for a variety of critical factors including gender, sampling methods, and occupation. 

In general, the mean serum levels following occupational exposures appear to be orders of

magnitude higher than observed in the general population. Thus, MOEs for workers would be

expected to be much less than for the general population.

5.3 Calculation of MOEs Based on Non-Human Primate Studies

As described in section 5.1, two studies have been conducted in non-human primates. A LOAEL

of 3 mg/kg-day was observed in both studies, and neither study established a NOAEL. For

calculation of MOEs, the 6-month male cynomolgus monkey study was chosen since it was a

more robust study. The effects observed at 3 mg/kg-day included increased liver weight and

possibly mortality. Serum levels of PFOA were measured throughout the study and a steady state

level of 77 + 39 :g/ml was observed in the 3 mg/kg-day group.

The most relevant general human population for use in the calculation of the MOEs is the

population of adults, age 20-69, that was examined in the American Red Cross samples (Table

27). Although the cynomolgus monkey study was conducted in males, it should be noted that

there is no evidence of a gender difference in the half-life of PFOA in humans or non-human

primates. Comparing males and females, there was no evidence of a gender difference in toxicity

in the Rhesus monkey study. In addition, the serum levels in the adult male and female human

populations are very similar. Therefore, MOEs were calculated using the serum levels from the

combined males and females. The MOEs were calculated by dividing the steady state serum

value for the male cynomolgus monkeys in the 3 mg/kg-day group (77 :g/ml) by the adult serum

levels from the American Red Cross blood samples presented in Table 28. The MOE calculated

using the geometric mean (.0046 :g/ml) for the human serum level is 16,739. The MOE

calculated using the 90th percentile value (.0094 :g/ml) for the human serum data is 8191.

5.4 Calculation of MOEs Based on Adult Rat Studies

As described in section 5.1, the effect levels that are appropriate for calculation of the MOEs are

the LOAEL of 1 mg/kg-day for adult male rats from the two-generation reproductive toxicity

study, and the NOAEL of 1.6 mg/kg-day for adult female rats from the 2-year study. Information

on the serum levels of PFOA were not available for either study. Therefore, pharmacokinetic

information was used to estimate the Area Under Concentration curve in plasma (AUC) as

described in Appendix A. The estimated AUC for adult male rats exposed to 1 mg/kg-day is1011 

:g-hr/mL and the estimated AUC for adult female rats exposed to 1.6 mg/kg-day is 44 :g-hr/mL.

The most relevant general human population for use in the calculation of the MOEs is the

population of adults, age 20-69, that was examined in the American Red Cross samples (Table

28). Assuming that the serum levels reflect steady state in humans, an AUC was calculated by

multiplying the serum concentrations by 24 hours. Thus, the AUC for the geometric mean is

0.1104 :g-hr/mL (0.0046 :g/mL X 24 hours), and the AUC for the 90th percentile is 0.2256 :ghr/mL (0.0094 :g/mL X 24 hours). MOEs were calculated by dividing the AUC in the adult

female rat by the AUC for the adult humans which is 398 (195 for the 90th percentile) and by

dividing the AUC for the adult male rat by the AUC for the adult humans which is 9158 (4481 for

the 90th percentile). 
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99

5.5 Calculation of MOEs Based on Rat Developmental Toxicity Studies

As described in section 5.1, developmental effects were observed in a two-generation

reproductive toxicity study in Sprague-Dawley rats (York, 2002). These effects were observed at

various times during the maturation of the F1 pups. During lactation, there was a reduction in F1

mean body weight on a litter basis (sexes combined) in the 30 mg/kg-day group. There was a

significant increase in mortality mainly during the first few days after weaning, and a significant

delay in the timing of sexual maturation for F1 males and females in the 30 mg/kg-day group. 

Mean body weights were also significantly reduced prior to sexual maturation in the F1 males

exposed to 10 and 30 mg/kg-day APFO and in the F1 females exposed to 30 mg/kg-day APFO.

The critical period of exposure for each of the effects is not known. For example, it is not known

whether prenatal and/or lactational exposure is important for the reduced body weight that was

observed during lactation. Similarly, it is not known whether the reduced body weight, mortality,

or delayed sexual maturation that occurred during the postweaning period are due to prenatal,

lactational, and/or postweaning exposures. Therefore, it is necessary to calculate MOEs for each

of these exposure periods.

Only limited information on serum levels of PFOA were obtained in the two-generation

reproductive toxicity study. This included measurements of serum levels in the F0 dams in the 10

and 30 mg/kg-day groups which were taken 24 hours after dosing. Given the rapid excretion of

PFOA by female rats, these values are of limited utility. In addition, serum levels were not

obtained for the F1 animals. Therefore, it was necessary to estimate serum levels based on

pharmacokinetic information. 

For the prenatal period, information was available on serum levels in the pregnant rat and fetus,

as well as concentrations of PFOA in the embryo. However, the information on serum levels in

the general U.S. population is restricted to children and adults; there is no information on levels in

the human embryo or fetus. Therefore, it is only appropriate to derive MOEs based on serum

levels in the pregnant rat and human females. Estimates of Cmax, and AUC in the pregnant rat

were used for calculation of the MOEs since it is not known which dose metric may be more

significant. These dose metrics were estimated for a pregnant rat administered 3 mg/kg-day

APFO since the lowest NOAEL for the developmental effects in the two-generation reproductive

toxicity study was 3 mg/kg-day. 

As described in section 3.2.4.1, PFOA serum levels were measured in the pregnant rat on

gestation days 10, 15 and 21 following exposure to 3, 10, or 30 mg/kg-day APFO (Mylchreest,

2003). The serum levels were determined two hours after dosing. Kemper (2003) determined

that peak serum concentrations of PFOA occur approximately 0.5 to 2 hrs after gavage dosing in

the adult female rat. Thus, the serum levels obtained by Mylchreest (2003) are approximately

equivalent to Cmax. An average Cmax for the pregnant rat was then calculated by averaging the

serum levels obtained at gestation days 10, 15 and 21; this value is 13 :g/ml for the 3 mg/kg-day

dose group. The AUC was estimated using a one compartment model which is described in

Appendix A. The AUC for the pregnant rat administered 3 mg/kg-day is approximately 83 :ghr/mL.

The most appropriate human population for comparison is the group of adult females, ages 20 -

69 years (Table 28). MOEs for the prenatal period were calculated by dividing the average Cmax

for the pregnant rat by the measured serum levels for adult human females which is 3,095 (1548

using the 90th percentile for human females), and by dividing the estimated AUC for the pregnant
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100

rat by the AUC for adult human females which is 823 (412 using the 90th percentile for human

females). 

Estimation of meaningful serum levels in the F1 pups during lactation is problematic. As

described in section 3.2.4.1, Mylchreest (2003) measured serum PFOA concentrations on days 3,

7, 14, and 21 of lactation. These measurements were taken approximately 2 hours after the dams

were dosed. Pups were separated from the dams 1-2 hours prior to the time that blood samples

were collected, and therefore it is not clear when the pups last nursed relative to the time that their

blood samples were collected. Han (2003) has shown that the clearance in 4-week old male and

female pups is fairly rapid. Assuming that the clearance is also rapid during the lactation period,

steady state would not be achieved. Thus, the measured serum concentrations in the pups would

not be expected to estimate steady state concentrations, and in fact, it is not clear what dose

metric the serum levels represent. The AUC for the pups could also be estimated, but, this

requires knowledge of milk consumption and elimination constants during lactation. While

estimates of milk consumption and elimination constants could be made, such estimates would

introduce considerable uncertainty in the resulting estimates of the AUC. Therefore, it is

premature at this point to derive estimates of the AUC for the lactation period. In addition, there

is no information from the available human biomonitoring studies on serum levels in children less

than two years of age. Consequently, MOEs were not calculated for the lactation period.

MOEs were calculated for the postweaning period. As described in section 3.2.4.2, Han (2003)

measured serum concentrations of PFOA in rat weanlings following a single dose of 10 mg/kgday APFO at 4, 5, 6, 7 and 8 weeks of age. Using a one compartment model, this information

was used to provide estimates of AUC (Appendix A). In the two-generation reproductive toxicity

study, several effects were noted during the postweaning period. One of these effects was

mortality that occurred mainly during the first few days after weaning in the male and female

groups exposed to 30 mg/kg-day APFO. For this effect, the most appropriate dose metric is the

estimated AUC for the 4-week old rat exposed to 30 mg/kg-day APFO; these estimates are 2,022

and 1,383 :g-hr/mL for the males and females, respectively. 

The most appropriate human population for comparison is the children of ages 2-12 (Table 28). 

The AUC for human children calculated from the geometric mean is 0.1176 :g-hr/mL (0.0049

:g/mL X 24 hours), and the AUC for the 90th percentile is 0.204 :g-hr/mL (0.0085 :g/mL X 24

hours). MOEs were calculated using the gender specific estimates for the AUCs for the 4-week

rat weanlings and the AUC for children. The MOEs calculated from the AUCs for male and

female rat weanlings, respectively, and the AUC for the geometric mean for children are 17,194

and 11,760, respectively. If the AUC for the 90th percentile for the human children is used, the

resulting MOEs are 9,912 and 6,779, respectively.

Another effect that was observed in the two-generation reproductive toxicity study was a delay in

the sexual maturation of the F1 females exposed to 30 mg/kg-day APFO. Mean body weights

were also reduced in this group prior to sexual maturation. Since sexual maturation occurs at

approximately 5 weeks of age, the most appropriate estimate of serum concentrations is the

average of the estimated AUCs for the 4 and 5 week female weanling exposed to 30 mg/kg-day

(Appendix A) which is 1233 :g-hr/mL. The MOE calculated using the average AUC for the 4

and 5 week female weanling rat and the AUC for the geometric mean for human children is

10,485 (6,044 using the AUC for the 90th percentile).

Sexual maturation was also delayed in the F1 males in the two-generation reproductive toxicity

study exposed to 30 mg/kg-day APFO. In addition, mean body weights were reduced prior to
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101

sexual maturation in F1 males exposed to 10 and 30 mg/kg-day. Sexual maturation occurs at

approximately 7 - 7.5 weeks, and the most appropriate estimate of serum concentrations is the

average of the estimated AUCs for the 4 - 8 week male weanling (Appendix A). The average

AUC at 10 mg/kg-day is 9237 :g-hr/mL. The MOE calculated from the average AUC for the 4 -

8 week male weanling rat and the AUC for human children is 78,546 (45,279 using the AUC for

the 90th percentile).

5.6 Uncertainties in the Risk Characterization

This assessment relies on a rich hazard data base for an industrial chemical. Extensive

pharmacokinetic information is available for rats through several life stages, and more limited

information is available for humans and non-human primates. Similarly, there are animal

toxicology studies for a variety of species including non-human primates, rats, mice and rabbits,

and more limited epidemiological studies. Extensive mode of action information is available that

has established that the liver toxicity and liver tumors are due to PPAR"-agonism. In addition,

serum biomonitoring information is available for humans throughout most of the life stages. 

PFOA serum levels were also obtained in some of the animal toxicology studies, and

pharmacokinetic data permitted the estimation of serum levels in some cases where measured

values were not available. These data permitted an analysis of potential human health effects

associated with exposure to PFOA utilizing a variety of animal models, several life stages, and

fairly precise estimates of actual human exposure.

Nonetheless, there are several uncertainties in this assessment. These are discussed below in a

qualitative fashion, since it is premature to attempt a quantitative analysis at this time. One area

of uncertainty pertains to the potential carcinogenesis associated with APFO exposure. In rats,

chronic exposure to APFO results in liver adenomas, Leydig cell adenomas, and pancreatic acinar

cell tumors. An extensive mode of action analysis has demonstrated that the liver tumors result

from PPAR"-agonism. However, there are insufficient information to establish the mode of

action for the Leydig cell and pancreatic acinar cell tumors. In addition, there is some uncertainty

regarding the development of mammary fibroadenomas following chronic APFO exposure in

rats.

In addition, it is not known which animal species is the most appropriate model for humans. The

prenatal developmental toxicity study in New Zealand White rabbits and the studies of

immunotoxicity in mice were not considered in the risk assessment since serum levels of PFOA

were not measured and there were no pharmacokinetic studies available with which to estimate

serum levels. In addition, the female rat may not be a good model for humans due to the gender

specific differences in elimination in rats. Information that is available for humans indicates that

there is no gender difference in elimination. Thus, human females would be expected to be

presented with a fairly continuous internal exposure to PFOA due to the long half-life, whereas

the female rat experiences an internal exposure that is rapid and discontinuous because of the

short half-life. This may be of particular importance in delineating the potential hazards

associated with exposure to PFOA during development. Indeed, the mouse does not appear to

have a gender difference in the elimination of PFOA, and preliminary results of a developmental

toxicity study in CD-1 mice indicate early pregnancy loss, compromised postnatal survival and

delays in postnatal growth and maturation following prenatal exposure to PFOA (Lau et al.,

2004).

Non-human primate data were used to assess the potential health effects to adult humans. There

is some uncertainty associated with the two studies that have been conducted due to the small

sample size and the lack of a NOAEL. There is also some uncertainty associated with the exact
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102

effects in the cynomolgus monkey study as it is not clear whether treatment-related mortality

occurred in the lower dose groups. Finally, there is some uncertainty regarding the serum levels

of PFOA in the cynomolgus monkey study due to the lack of a linear increase in serum PFOA

levels in proportion to dose, the high variability, and the precision of the method itself which was

+ 30% (inter-assay, intra-assay, and system).

Data from several toxicology studies in adult rats were also used to estimate the potential for

health effects in adult humans. The only consistent effect that was noted in adult animals was a

reduction in body weight. The implications of this effect for human health are not clear, but the

endpoint was used to provide a conservative benchmark from a second species. As noted above,

the female rat may not be a good model for human females. It is possible, due to the kinetics in

the female rat, that some effects were not observed in the rat studies that may be observed in

species with a longer half-life. In addition, there was a dose-related increase in the incidence of

ovarian tubular hyperplasia in the 2-year study in rats. This effect was not observed in studies of

shorter duration and did not progress to tumor formation in the 2-year study. Thus, the

significance of this finding is not clear. Further research will need to be conducted to determine

the significance of the finding for humans.

Several developmental effects associated with PFOA exposure were observed in the twogeneration reproductive toxicity study in rats. The study did not determine the critical period of

exposure for the developmental effects and serum levels were not measured. Therefore, it was

necessary to use pharmacokinetic data to estimate the serum levels for the prenatal, lactation, and

post-weaning periods. For the prenatal period, dose metrics were calculated for the pregnant

dams; dose metrics were not calculated for the embryo/fetus since comparable dose metrics do

not exist for humans. As described in Appendix A, pharmacokinetic data in adult female rats

were used to estimate values for the volume of distribution, and elimination and absorption rate

constants; this information was then used to estimate the AUC in the pregnant rat. Although one

study did not report a difference in the elimination of PFOA during pregnancy in the rat (Gibson

and Johnson, 1983), there is still some uncertainty in whether the kinetics are the same in the

pregnant and non-pregnant rat. There is also some uncertainty whether the distribution of PFOA

to the embryo/fetus is similar in rats and humans. 

Unfortunately, dose metrics during the lactation period were not calculated. Therefore, it is not

known what the potential MOE may be for this exposure period. Methods to estimate the dose to

the offspring during lactation are under development.

Dose metrics for the post-weaning period were calculated. As described in Appendix A, data for

post-weaning rats permitted estimation of the elimination constants during the post-weaning

period. The adult volume of distribution and absorption constants were used; thus, there is some

uncertainty whether the volume of distribution and absorption constants apply across all life

stages. 

The pharmacokinetic modeling described in Appendix A considered both one and two

compartment models. Given the substantial database created for single dose pharmacokinetic

studies in the rat, there would be expected to be relatively small uncertainties estimating AUCs. 

In general the fits to the plasma time course data were not substantially improved with the two

compartment model as indicated by generally similar AIC values (see Tables A-xii and A-xv). 

Given the increased number of parameters in the two compartment model AIC values would have

to be substantially lower for that model to give a statistically improved fit. Visual inspection

showed that the divergence in plasma time course from the one compartment fit tended to occur

at later times in some dose groups. With the females, it was impossible to obtain two-
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103

compartment fits at lower concentrations due to rapid clearance and analytical detection limits. 

Given the relatively short period of dosing for the reproductive/developmental studies, error

introduced through use of the one compartment model would be relatively small. For the male

animals, doses that appeared well fitted by one compartment were interspersed with doses

suggestive of a second compartment. Use of the one compartment model to estimate AUCs for

chronic duration studies may tend to introduce greater error, though the greater uncertainty for the

chronic studies is whether extended dosing alters physiology and thus pharmacokinetics.

There is also some uncertainty regarding the use of the human biomonitoring data. Although the

available data include a range of populations with various demographic features, there may be

some populations that are not represented. Since it is unknown how the human exposures are

occurring, proximity to a manufacturing plant may be a factor in exposure. However, populations

living near the plants were not sampled. Therefore, it is possible that PFOA serum levels may be

underestimated for certain portions of the U.S. population. The children’s sample was derived

from blood collected in 1994/1995; therefore, it may not reflect the current status of PFOA in

children’s blood. It is not clear how PFOA may affect more sensitive subpopulations or if their

exposures would vary. PFOA will be included in future environmental reports of the National

Health and Nutrition Examination Survey (NHANES), in an effort to acquire nationally

representative serum PFOA levels; however, children's serum levels will not be included. 

Finally, there is some uncertainty in the interpretation of the MOEs. MOEs are generally

calculated by dividing an administered dose in an animal study by an estimate of external human

exposure. The value of the MOE that is associated with a concern for toxic effects is generally

expressed as the product of the applicable uncertainty factors. The uncertainty factors that are

considered are the same as those considered in the derivation of a RfD or RfC. Generally this

includes consideration of an uncertainty factor for intraspecies variation and another for

interspecies variation. An additional factor may be considered if a LOAEL is used from the

animal toxicology study. In this assessment, MOEs were calculated from serum levels in animals

and humans. Generally, data are not available for this kind of estimate, and therefore there is

little experience in determining whether the same uncertainty factors apply. In addition, this

assessment is unique in that the half-life of PFOA is very different among the various animal

species and humans. Whether this should be a consideration in interpreting the MOE is not clear.

6.0 Overall Conclusions

This risk assessment focused on the potential human health effects associated with exposure to

PFOA and its salts. The results of existing epidemiology studies are not adequate for use in

quantitative risk assessment, and therefore the analysis was restricted to endpoints in the animal

toxicology studies. MOEs were calculated for the general U.S. population. Although some

serum level data were available for workers, the data were not adequate to calculate MOEs for

specific occupational exposures. In general, the mean serum levels following occupational

exposures appear to be orders of magnitude higher than observed in the general population. Thus,

MOEs for workers would be expected to be much less than for the general population.

A variety of endpoints from the animal toxicology studies were used to calculate MOEs for this

draft risk assessment. The endpoints encompassed different species, gender and life stages. For

this draft assessment, specific recommendations on the most appropriate

endpoint/lifestage/species/gender have not been made; rather, all have been presented to provide

transparency.
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104

For adults, two sets of MOEs were calculated based on the toxicology studies in non-human

primates and rats. First, calculation of MOEs from the cynomolgus monkey study was based on

increased liver weight and possible mortality. The MOE using the geometric mean for the human

serum level is 16,739 (8,191 for the 90th percentile). Second, calculation of MOEs from the adult

rat studies was based on reductions in body weight. MOEs were calculated separately for the

female and male rat due to the gender differences in pharmacokinetics in this species. MOEs

were calculated by dividing the AUC in the adult female rat by the AUC for the adult humans

which is 398 (195 for the 90th percentile) and by dividing the AUC for the adult male rat by the

AUC for the adult humans which is 9158 (4481 for the 90th percentile).

MOEs were calculated for the developmental effects in the two-generation reproductive toxicity

study in rats. These effects were observed at various times during the maturation of the F1 pups;

the critical period of exposure for each of the effects is not known, and could be due to exposure

during the prenatal, lactational, and/or postweaning periods. Ideally, MOEs should be calculated

for each of these exposure periods; however MOEs were not calculated for the lactation period

due to uncertainties in pharmacokinetics. For the prenatal period, MOEs were calculated for the

pregnant human female; MOEs were not calculated for the fetus since there is no information on

human serum levels in fetuses. MOEs were calculated using both Cmax and AUC; the MOE based

on Cmax is 3,095 (1548 for the 90th percentile) and the MOE based on the AUC is 823 (412 for the

90th percentile). 

For the postweaning period, MOEs were calculated for several endpoints including reductions in

body weight, mortality and delayed sexual maturation. These MOEs were based on the geometric

mean for children and range from 10,484 - 78,546 (the range using the 90th percentile is 6,044 -

45,279).

This assessment has provided a range of MOEs for several life stages. Several uncertainties have

been discussed in a qualitative fashion in this assessment, which highlight the need to interpret

the MOEs with caution. For example, MOEs were not calculated for the lactation period due to

insufficient data, and this life stage may represent an important exposure period. Similarly, the

biomonitoring data for the children are from samples collected in 1994 and may not be

representative of current children’s serum levels. In addition, ascertaining potential levels of

concern will necessitate a better understanding of the appropriate dose metric in rats or other

animal models, and the relationship of the dose metric to the human serum levels. 

Finally, there is some uncertainty associated with the determination of the adequacy of a specific

MOE in protecting human health in the present context. Generally, MOEs are calculated from

administered dose levels and estimates of human exposure. In this assessment, the MOEs were

calculated from internal dose metrics in animals and humans. While use of internal dose metrics

reduces many uncertainties pertaining to exposure, there is little experience or guidance on the

factors that should be considered in making judgements about the level of concern associated

with a given MOE. Approaches that are used for conventional MOEs, if applied unchanged,

indicate that among the populations of interest some individuals are highly exposed, for reasons

not understood at this time. However, if conventional approaches for determining levels of

concern are not appropriate for MOEs based on internal dose metrics, then this conclusion would

have to be re-evaluated as the understanding of this question evolves. 
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105

7.0 References

3M (Minnesota Mining and Manufacturing Company). 1999a Letter (dated May 26, 1999) re:

TSCA 8(e) Supplemental Notice: Sulfonate-based and carboxylic-based fluorochemicals–Docket

Nos. 8EHQ-1180-373; 8EHQ-1180-374; 8EHQ-0381-0394; 8EHQ-0699-373.

3M (Minnesota Mining and Manufacturing Company). 1999b The Science of Organic

Fluorochemistry” and “Perfluorooctane sulfonate: current summary of human sera, health, and

toxicology data.” February 5, 1999. 8EHQ-0299-373.

3M (Minnesota Mining and Manufacturing Company). 2001 Material Safety Data Sheet FC-26

FLUORAD Brand Fluorochemical Acid, ID Number/U.P.C.: ZF-0002-0376-8. Online: 

http://www.3m.com/US/safety/index.jhtml. 

Abdellatif, A.G.; Preat, V.; Taper, H.S. and Roberfroid, M. 1991 The modulation of rat liver

carcinogenesis by perfluorooctanoic acid, a peroxisome proliferator. Toxicol. Appl. Pharmacol.

111:530-537.

Alexander, B.H. 2001a Mortality study of workers employed at the 3M Cottage Grove facility. 

Final Report. April 26, 2001. Division of Environmental and Occupational Health, School of

Public Health, University of Minnesota. US EPA AR226-1030a018.

Alexander, B.H. 2001b Mortality study of workers employed at the 3M Decatur facility. Final

Report. April 26, 2001. Division of Environmental and Occupational Health, School of Public

Health, University of Minnesota. US EPA AR226-1030a019.

Berthiaume, J. and Wallace, K.B. 2002 Perfluorooctanoate, perfluorooctanesulfonate, and Nethyl perfluorooctanesulfonamido ethanol; peroxisome proliferation and mitochondrial

biogenesis. Toxicol. Lett. 129:23-32.

Biegel, L.B.; Liu, R.C.M; Hurtt, M.E. and Cook, J.C. 1995 Effects of ammonium

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A-1

Appendix A: Predicting serum concentrations in rats

A-1.0. Introduction

Analysis of pharmacokinetics for PFOA could be done using a number of approaches including

physiologically based pharmacokinetic (PBPK) modeling, classical compartmental modeling, and

non-parametric analysis. Each of these has strengths and limitation given the available data. 

Some of the literature and submitted pharmacokinetic studies have been analyzed using nonparameteric analyses. This provides a description of the data that have been collected, but has

fairly limited ability to make predictions across species or variations in exposures. Classical

compartmental modeling can be used to analyze the existing data on blood concentrations in rats,

monkeys, and humans. Much of the published analyses use this approach and the data in humans

and perhaps monkeys is only adequate to support this approach. Comparisons of serum protein

binding across species indicated a high degree of binding in all species eliminating the apparent

need to address this factor in the compartment modeling. PBPK modeling is perhaps the ideal

approach for addressing PFOA for purposes of cross-species extrapolation. Extensive

pharmacokinetic studies have been undertaken in rodents demonstrating complex phenomenon

such as high tissue concentrations in liver, kidney and serum and enterohepatic recirculation of

the parent compound. These could be addressed using PBPK modeling for the rodents, but the

more limited information in monkeys and humans would either require substantial assumptions or

preclude use of this approach. In light of the documented differences in clearance of PFOA

across sexes in rats and across species, compartmental modeling of serum concentrations

provides a sound approach for estimating internal dosimetry without exceeding the limits of the

available data, so this approach was selected for this risk assessment.

The calculations (see Sections 5.4 and 5.5) of margins of exposures (MOEs) using human and

rodent plasma concentrations or the Area Under Concentration curve in plasma (AUC) require

either direct measurements or prediction of these concentrations and AUCs. AUC is a measure of

dose that reflects both concentration and time; i.e., it increases with higher concentrations and

longer durations. The AUCs estimated here are the average daily AUC under conditions of

repeated dosing. Concentrations in plasma at the time of measurement reflect the time course for

circulating chemical including the maximum levels achieved following a bolus dose; with slow

clearance and repeated dosing they can reflect steady state conditions. Pharmacokinetic studies in

Sprague-Dawley rats (Kemper (2003); Han (2003); see Section 3.2) provided information for

parameterizing and calibrating a predictive pharmacokinetic model using standard software

(WinNonlin® Version 4.0.1). Measurements of plasma concentrations in several other studies

with Sprague-Dawley rat (York, 2002; Mylchreest, 2003; Palazzolo, 1993; Goldenthal, 1978a)

provide data for evaluating the model (see Section A-4.0) and in some cases for direct

comparison to human measured plasma concentrations in Section 5.5.

Compartmental models were fitted to the data for individual adult animals in Kemper (2003) (see

Section A-5.0). WinNonlin® Version 4.0.1 was applied to oral plasma time course data to

estimate and compare one-compartment and two-compartment models with both absorption and

elimination constants. Similar fitting was done for the intravenous plasma time course data. 

Akaike Information Criteria and correlations of observed and predicted values at each dose

holistically afforded a basis for the comparisons of the fits and the one- and two-compartment

models. For either sex, in the range of dosing of most interest, there was little difference among

the models, although there is a large difference in elimination rate between adult male and female

rats.
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A-2

A one-compartment model was selected for predicting plasma concentrations and AUCs due to its

ease of use and consistency with the data. The plasma concentration in such a model with firstorder oral absorption and excretion can be expressed (see O’Flaherty, 1981, equation 3.42) as:

 C

k D

Vk k

e e a

a e

kt kt e a = − − − −

( ) ( ).

where C is the plasma concentration at a given time (mg/L), D is the dose (mg), V is the volume

of distribution (L), ka is the absorption rate constant (/hr), ke is the elimination rate constant (/hr),

and t is time (hr). Then AUC = D/(Vke); this is expressed in mg-hr/L or equivalently, ug-hr/mL. 

Sex-specific averages of V, ke and ka were calculated across the one-compartment model fits to

all the orally-administered doses. Further, AUC can be expressed in terms of the frequently

reported body-weight-adjusted values for dose and volume of distribution. D can be expressed as

D = body weight x *, where * is the administered dose rate in mg/kg per day, and V can be

written V = body weight x $, where $ is a per-unit-weight value in L/kg. Thus, AUC = */($ke),

also in mg-hr/L or ug-hr/L.

Table A-i Pharmacokinetic Model Values for Adult Rats

$ (L/kg) ka (/hr) ke (/hr) t1/2 (hr)

Male Rat 0.172 3.87 0.0058 119

Female Rat 0.168 3.73 0.216 3.2

Estimates of AUC were determined for the doses, exposure durations, and life stages relevant to

the toxicity studies for which calculations were planned for Section 5. For gavage dose studies,

the oral dose received by the animals was known. For dietary studies, the daily dose was

estimated based upon food consumption and chemical concentration in the diet. AUCs using the

parameter values as described above were predicted for toxicity studies involving adult male and

nonpregnant female rats. The two-generation reproductive toxicity study includes males,

pregnant females, lactating females, developing fetuses, lactationally exposed offspring, and

juvenile offspring, so estimates of AUC are desirable for each of these lifestages. Estimates of

AUC for gavage-dosed males, pregnant females, and lactating females used the parameter values

as described above. As indicated in Section A-4.0, plasma data from pregnant and lactating

females were reasonably predicted by the modeling without adjustments for these lifestages. 

Owing to the evidence in the Han (2004) study with juvenile animals that the clearance in males

drops dramatically with sexual maturation after week 4, these data were used to re-estimate

weekly values of ke for gavage dosed juvenile males and females. No estimates for AUC for

lactationally-exposed offspring were developed owing to uncertainties in the dose they received,

as discussed in Section 5.

A-2.0. Adult Male and Female Rats

A-2.1. Adult Male Rats

Based on Kemper (2003), adult male rats have relatively slow clearance, with average half life

just under a week (about 5 days). It takes 21 days to be within 95% of steady state levels. 

Consequently, adult male rats in toxicity studies longer than 21 days would be expected to be at

steady state, though neither the single dose nor the extended course pharmacokinetic studies

would have achieved steady state. Data from the subchronic toxicity study by Palazzolo (1993),
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A-3

which extended 14 weeks, can be used to evaluate the adequacy of predictions based upon single

doses for repeated dosing of male rats to steady state (see Section A-4.0).

At steady state, 

Csteady state or CSS = (dose rate mg/kg)/((male ke /hour)*(male $ L/kg)*(24hour/day))

AUC= 24*(conc at steady state) = 24*CSS

Table A-ii.a. Values for Adult Male Rats

administered dose

(mg/kg)

1 3 10 30

AUC (:g-hr/mL) 1011 3032 10108 30324

CSS (mg/L) 42 126 421 1263

A-2.2. Adult Female Rats

Females have rapid clearance, with an estimated half-life of 0.134 days or 3.2 hours. The female

will clear virtually the entire gavage or dietary dose each day, so plasma concentrations will vary

throughout the day, but successive days will appear similar. The process does not reach steady

state, but an AUC can be estimated.

Table A-ii.b. Values for Adult Female Rats

administered dose

(mg/kg)

1.6 3 10 30

AUC (:g-hr/mL) 44.16 82.80 276.01 828.03

A-3.0. Post-weaning, Juvenile Rats

To estimate AUCs for the gavage dosed juvenile rats in the two-generation toxicity study, the

data from Han (2003) were used to estimate ke in a week- and sex-specific manner. Supposing

the adult per-unit-weight volume of distribution $ estimated from Kemper (2003) holds for the

offspring, the elimination constant can be obtained from the Han (2003) data using

 C(T)=(D/V)(exp(-keT))=(*/$)(exp(-keT))

where C(T) is the concentration at 24 hours after gavage dosing, * = 10 mg/kg as used in Han

(2003), and $ is the adult sex-specific weight-adjusted volume of distribution from the

WinNonlin® solutions to Kemper (2003). Absorption is long completed by 24 hours, so an

instantaneous absorption assumption is not very influential. That is, the concentration equation

above is essentially the same as the one in Section A-1.0, when ka is large relative to ke. In

Section A-5.0 it will be seen ka/ke is about three orders of magnitude for male adult rats and one

and a half for female adults. The impact of their relative size has been further verified for the rats

by successfully re-estimating the 24-hour plasma concentrations of Han (2003) using the

expanded equation, the weekly elimination rates, and the adult absorption rates (results not

shown).
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A-4

Then,

 ke /hr = (-loge($*C(T))/*)/T 

 = (-ln($ L/kg*(C(24 hr) mg/L)))/* mg/kg)/(24 hr)

This sex-specific ke is then used with the equation AUC = D/(Vke)= (*/($ke)) to obtain the AUCs. 

A-3.1 Post-weaning, Juvenile Male Rats

For the males, half lives range from a third of a day to 2.5 days throughout the post-weaning

period of four to eight weeks of age in Han (2003). The longer half-life observed from weeks 5

to 8 would require 10 days to be within 95% of steady state. Treating the juvenile males as if

they are at steady state is, thus, an approximation till steady state would be achieved around week

6 or 7.

 

 AUC = 24 hr * δ mg/kg/day /(β L/kg * ke /hr * 24 hr/day)

Table A-iii. Post-weaning Male Rats, AUC, using 0.172 = adult male $

 dose rate 1 3 10 30

4wk AUC (:g-hr/mL) 67.40 202.19 673.96 2021.87

5wk AUC (:g-hr/mL) 354.25 1062.76 3542.53 10627.59

6wk AUC (:g-hr/mL) 468.13 1404.39 4681.29 14043.85

7wk AUC (:g-hr/mL) 310.66 931.98 3106.62 9319.84

8wk AUC (:g-hr/mL) 339.11 1017.32 3391.07 10173.20

mean wks4-8 307.91 923.73 3079.09 9237.27

A-3.2. Post-weaning, Juvenile Female Rats

For the females, ke estimated in a week-specific manner from Han (2003) can also be applied. A

sex-specific ke is again used with the equation AUC = D/(Vke)= (*/($ke)) to obtain the AUCs. 

Half-lives of the female pups range about 3.5-5.5 hours, so steady state would not be achieved

following gavage dosing.

 AUC = δ mg/kg / (β L/kg * ke /hr)

Table A-iv. Post-weaning Female Rats, AUC, using 0.168 = adult female $

dose (daily) 1 3 10 30

4wk AUC (:g-hr/mL) 46.11 138.33 461.11 1383.34

5wk AUC (:g-hr/mL) 36.07 108.21 360.69 1082.06

6wk AUC (:g-hr/mL) 36.47 109.40 364.67 1094.01

7wk AUC (:g-hr/mL) 30.76 92.28 307.60 922.81

8wk AUC (:g-hr/mL) 33.28 99.82 332.75 998.25

mean wks4-5 41.09 123.27 410.90 1232.70

A-4.0. Validation of One-compartment Model for Predicting Dose Metrics

The parameters for the one-compartment pharmacokinetic model applied to PFOA were

derived from fits using WinNonlin®4.0.1 with single oral dose pharmacokinetic studies in male

and female Sprague-Dawley rats (Kemper 2003). These studies measuring plasma time courses

used doses of 0.1, 1 (two different studies with different durations), 5, and 25 mg/kg. The

average parameter values across these five data sets were used when making predictions. A
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A-5

single intravenous dose (1 mg/kg) in male and female Sprague-Dawley rats provides independent

estimates of the volume of distribution and elimination rate that are reasonably consistent with the

average values from the oral studies (Kemper, 2003). Another Kemper (2003) study in male and

female Sprague-Dawley rats showed no major changes in the extent of total urinary and fecal

elimination for a 1 mg/kg oral dose given following 14 days of dosing at 1 mg/kg/day (i.e., there

were no apparent pharmacokinetic changes due to repeated dosing; see Section 3.2.3.5.2.1). 

Additional urinary and fecal elimination studies following single doses are available, but have not

yet been subjected to quantitative analysis using the one-compartment model.

The parameters for males and females can be used to predict blood plasma concentrations

that were measured in several of the toxicity studies. Plasma levels following gavage dosing also

have been reported for adult female rats in the tissue distribution studies of pregnant rats and

lactationally exposed offspring (Mylchreest, 2003). Parental male and female Sprague-Dawley

rats in the 2-generation study had blood sampled at the end of mating and lactation, respectively

(York, 2002). Two dietary studies exposed rats for 90 days at the end of which blood samples

were obtained. One study used male Sprague-Dawley rats (Palazzolo, 1993). The other used

male and female Sprague-Dawley rats (Goldenthal, 1978a).

Different analyses are required for predicting the plasma concentrations in these studies

using the one-comparment model. Owing to rapid clearance in females, it is important to predict

the concentration at the appropriate time following dosing, even when it is measured in a repeated

dosing study. Modeling dietary exposure for females is likely to be somewhat problematic using

a simple bolus approach. For males, their relatively slow clearance would result in achieving

steady state plasma levels following sufficient duration of dosing; it should make little difference

whether the exposure was by gavage or diet.

The plasma concentrations in pregnant and lactating rats 2 hrs after dosing are reasonably

well predicted using the average values from the oral pharmacokinetic studies in females for the

volume of distribution, elimination rate, and oral absorption rate (Mylchreest, 2003). The

predicted concentrations at the higher doses exceed the measured levels by less than a factor of

two, while the lower dose is well predicted. Small differences in the kinetic parameters would be

expected for pregnant animals; pregnancy might lower their predicted plasma concentrations

(owing to a larger volume of distribution, reduced protein binding and increased clearance). 

Lactating females eliminate some fraction of their body burdens in their milk, which may also

reduce their plasma concentrations relative to these predictions.

Table A-v. Predicting Mylchreest (2003) 2-hour dam plasma concentrations (gavage)*

 Time

(hr)

Daily Dose

(mg/kg)

predicted C(2hr)

(ug/mL)

measured

(ug/mL)

measured SD

2 3 12.29 11.19 2.76 

10 40.96 26.84 4.21 

30 122.87 66.64 9.8 

 *C(2hr) calculated from 

 C t

k D

Vk k

e e

k

k k

e e a

a e

kt kt a

a e

kt kt ea ea ( ) ( ) ( ) ( ) = ( ). − − = − − −− −− δ

β

The plasma concentrations 24 hrs following gavage dosing of lactating female rats also

are reasonably well predicted at both the 10 and 30 mg/kg/day dose levels (York, 2002).
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A-6

C t

k D

Vk k

e e

k

k k

e e a

a e

kt kt a

a e

kt kt ea ea ( ) ( ) ( ) ( ) = ( ). − − = − − −− −− δ

β

Table A-vi. Predicting York (2002) 24-hr dam plasma concentrations (gavage)*

Time (hr) Daily Dose

(mg/kg)

predicted C(24 hr)

(ug/mL)

measured

(ug/mL)

measured SD

24 10 0.35 0.37 0.0805 

30 1.06 1.02 0.425 

 *C(24hr) calculated from 

Steady state plasma concentrations in male rats exposed via diet for 90 days were

predicted within a factor of two at a wide range of doses (Palazzolo, 1993). 

Table A-vii. Predicting Palazzolo (1993) male plasma concentrations (diet)*

 Time

(hr)

Diet

(ppm)

Dose Rate

(mg/kg/day)

Dose Rate SD

(mg/kg/day)

predicted Css

(ug/mL)

measured avg 5, 8, 14 wks

(ug/mL)

24 1 0.06 0.013 3 7.0 

10 0.64 0.138 27 47.4 

30 1.94 0.408 82 87.0 

100 6.5 1.214 274 148.7 

 *calculated as CSS (mg/L) = (dose rate mg/kg/day)/(($ L/kg)*(ke /hour)*(24hour/day))

These data and predicted levels provide confidence that the model using parameters

independently determined in the pharmacokinetic study is able to predict dose metrics in the

toxicity studies.

There are two groups of data that are not successfully predicted, the male plasma

concentrations from the York (2002) and Goldenthal (1978a) studies. It is important to note that

the measured concentrations in these studies are not consistent with the results of the Palazzolo

(1993) study, so it would be impossible for a single set of parameter values to predict all these

data sets. The source of this difference in measured results is unclear, particularly for the York

(2002) results since the female data from that study appear reasonably predicted. The analysis of

plasma from the Goldenthal (1978a) study involved considerably different analytical methods,

which may be an important factor.

The York (2002) male data reports plasma concentrations at 10 and 30 mg/kg/day that are

similar to those observed at 0.6 and 1.9 mg/kg/day (estimated for dietary exposure) in the

Palazzolo (1993) study. Treating the data as concentrations 24 hrs following a single gavage

gives a reasonable prediction for the lower dose, but not the higher one. Average concentrations

at 24 hr following a single oral dose of 25 mg/kg in Kemper (2003) are 127 ug/mL, clearly

inconsistent with the 45 ug/mL plasma level measured following repeated gavage in York (2002)

males. The males in York (2002) were being mated, but the potential for that to contribute to

changes in pharmacokinetics is unclear.
SAB Review Draft; Do Not Cite or Quote

A-7

Table A-viii. Predicting York (2002) 24-hr male plasma concentrations (gavage)*

Time (hr) Dose Rate

(mg/kg/day)

predicted Css

(ug/mL)

measured

(ug/mL)

measured SD predicted C(24hr)

(single dose)

(ug/mL)

24 10 421 51.1 9.3 50.77 

30 1263 45.3 12.6 152.31 

 *calculated as CSS (mg/L) = (DR mg/kg/day)/(($ L/kg)*(ke /hour)*(24hour/day)) and at 24 hr

from a single dose as 

 C t

k D

Vk k e e

k

k k e e a

a e

kt kt a

a e

kt kt ea ea ( ) ( ) ( ) ( ) = ( ). − − = − − −− −− δ

β

The plasma data in male rats exposed by diet for 90 days should similarly represent

pseudo-steady state concentrations (Goldenthal, 1978a). Plasma from these animals was analyzed

and reported separately by Belisle (1978). While the plasma levels at the lowest dietary exposure

are well predicted, the measured plasma values at higher dietary levels show little increase with

dose and are not well predicted.

Table A-ix. Predicting Goldenthal (1978a) male plasma concentrations (diet)*

Time (hr) Diet (ppm) Daily Dose

(mg/kg)

predicted Css (ug/mL) measured

(ug/mL)

24 10 0.6 25 21 

30 1.7 72 34 

100 5.6 236 36 

300 17.9 754 38 

1000 63.5 2674 49 

 *calculated as CSS (mg/L) = (dose rate mg/kg/day)/(($ L/kg)*(ke /hour)*(24hour/day))

The plasma data from female rats exposed by diet for 90 days are more difficult to predict

owing to the rapid clearance and the uncertainty in when the rats ate, so these data have not been

considered.
SAB Review Draft; Do Not Cite or Quote

A-8

TABLES

A-5.0. Results of the WinNonlin® Analyses
A-5.0. Results of the WinNonlin7

 Analyses

The WinNonlin7

 solutions corresponding to V , k a and k e are denoted in its output by V_F, 

K01, and K10, respectively, for the 1-compartment model (additional different coefficients are 

estimated for a 2-compartment model as shown in tables in A-5.0 for these analyses).

A-9
Table A-x. Female Rats, One-Compartment Model

Variable N Nmiss Nobs Mean SD SE Variance Min Median Max Range CV% Geometric MHarmonic MPseudo SD Mean Log SD Log 

0.1 mg/kg femaleV_F 3 0 3 28.31038 0.953683 0.550609 0.909511214 27.26332 28.5385 29.12932 1.865992 3.368669 28.29959 28.28871 0.970518 3.342847 0.03389K01 3 0 3 6.522584 1.525262 0.88061 2.326424021 4.821422 6.978249 7.768082 2.94666 23.38432 6.393595 6.257274 1.778985 1.855297 0.25022K10 3 0 3 0.190299 0.035546 0.020522 0.001263511 0.151857 0.19706 0.221978 0.070121 18.67902 0.187981 0.18559 0.038933 -1.671412 0.19416BW 4 176.4 8.6beta (L/kg) 0.160490.1 mg/kg female (extended time course)V_F 4 0 4 26.85968 5.712687 2.856344 32.63479772 21.29838 25.76741 34.60553 13.30716 21.26863 26.42605 26.01866 5.13963 3.27435 0.20626K01 4 0 4 1.731362 0.732875 0.366437 0.537105296 1.034098 1.717244 2.456863 1.422765 42.32936 1.61078 1.498008 0.656779 0.476718 0.44447K10 4 0 4 0.221401 0.042599 0.021299 0.001814668 0.172791 0.220222 0.272367 0.099576 19.24064 0.218303 0.215218 0.042948 -1.52187 0.19463BW 4 201 11.8beta (L/kg) 0.133631.0 mg/kg female intravenousV (mL) 4 0 4 44.7764 4.3264 2.1632 18.7175 41.0724 43.8474 50.3385 9.2661 9.6622 44.623 44.4733 4.1263 3.7982 0.095K10 (1/hr) 4 0 4 0.1703 0.029 0.0145 0.0008 0.1391 0.1713 0.1995 0.0603 17.0288 0.1684 0.1665 0.0291 -1.7814 0.172BW 4 200.6 2.7beta (L/kg) 0.2232121.0 mg/kg female oralV_F 4 0 4 35.77252 7.683779 3.841889 59.04045812 24.93616 37.81364 42.52666 17.59049 21.47955 35.07346 34.30106 9.588217 3.557445 0.23701K01 4 0 4 4.989439 2.068961 1.034481 4.280600063 2.599878 5.147627 7.062624 4.462746 41.46681 4.633671 4.272174 2.298085 1.533349 0.4592K10 4 0 4 0.161888 0.0214 0.0107 0.000457953 0.148536 0.152607 0.193803 0.045267 13.21891 0.160907 0.160005 0.018132 -1.826932 0.12493BW 4 197 9.7beta (L/kg) 0.1815865 mg/kg female oralV_F 4 0 4 31.74406 13.54995 6.774977 183.6012501 18.41039 30.6327 47.30047 28.89008 42.685 29.53732 27.48214 12.25893 3.385655 0.44301K01 4 0 4 2.152185 1.982879 0.99144 3.931809422 0.412166 2.072876 4.050822 3.638655 92.13331 1.302084 0.787231 0.794923 0.263966 1.25642K10 4 0 4 0.334105 0.121041 0.06052 0.014650842 0.230922 0.31729 0.470918 0.239996 36.22836 0.317682 0.302361 0.107012 -1.146704 0.36769BW 4 187.3 3.6beta (L/kg) 0.16948225 mg/kg female oralV_F 4 0 4 37.75743 9.235502 4.617751 85.29449457 24.25588 41.51663 43.74058 19.48471 24.46009 36.74818 35.58839 12.54934 3.604089 0.28140K01 4 0 4 3.941804 1.961216 0.980608 3.846368692 2.886762 2.99903 6.882396 3.995634 49.75427 3.65612 3.452334 1.045302 1.296402 0.42214K10 4 0 4 0.171864 0.009952 0.004976 9.90461E-05 0.159052 0.174024 0.180355 0.021303 5.79074 0.171644 0.171421 0.010261 -1.762331 0.05863BW 194.8 14.4beta (L/kg) 0.193827

A-10
Table A-xi. Female Rats, Two-Compartment Model

Variable N Nmiss Nobs Mean SD SE Variance Min Median Max Range CV%

Geometric Mean

Harmonic Mean

Pseudo SD Mean Log SD L

1 mg/kg oralV1_F 4 0 4 18.6234 21.3178 10.6589 454.4486 0.1737 16.9581 40.4038 40.2301 114.4676 3.2585 0.5067 1.4802 1.1813 2V2_F 4 0 4 9.0363 6.1941 3.0970 38.3667 2.0499 9.2846 15.5262 13.4762 68.5465 6.9706 4.9841 7.8707 1.9417 0K01 4 0 4 2.7225 2.7532 1.3766 7.5800 0.3420 2.2161 6.1158 5.7739 101.1273 1.4950 0.8091 1.2047 0.4021 1CL_F 4 0 4 6.0166 1.7649 0.8825 3.1150 3.9358 5.9444 8.2420 4.3062 29.3344 5.8175 5.6171 1.8777 1.7609 0CLD2_F 4 0 4 22.5062 15.8217 7.9109 250.3273 9.7486 17.3694 45.5376 35.7890 70.2993 19.0904 16.6218 10.1250 2.9492 05 mg/kg oralV1_F 3 0 3 29.4424 15.5140 8.9570 240.6835 18.7050 22.3928 47.2294 28.5244 52.6926 27.0454 25.1483 10.1302 3.2975 0V2_F 3 0 3 15.4949 20.3798 11.7663 415.3361 3.0245 4.4472 39.0132 35.9887 131.5255 8.0658 5.1624 4.3887 2.0876 1K01 3 0 3 1.6408 2.0681 1.1940 4.2771 0.4126 0.4814 4.0285 3.6160 126.0400 0.9283 0.6317 0.4303 -0.0743 1CL_F 3 0 3 9.3461 1.1661 0.6733 1.3598 8.4231 8.9586 10.6567 2.2336 12.4771 9.2992 9.2540 1.0768 2.2299 0CLD2_F 3 0 3 0.2948 0.0961 0.0555 0.0092 0.1958 0.3009 0.3877 0.1919 32.5940 0.2837 0.2725 0.1066 -1.2598 025 mg/kg oralV1_F 4 0 4 37.9011 8.5209 4.2605 72.6061 25.4681 41.1476 43.8412 18.3732 22.4820 37.0614 36.1080 11.2420 3.6126 0V2_F 4 0 4 81.1708 126.4525 63.2263 15990.2410 3.0616 26.0788 269.4641 266.4025 155.7857 25.4384 9.2358 30.9321 3.2363 1K01 4 0 4 13.3420 18.3460 9.1730 336.5755 2.7921 4.9360 40.7041 37.9120 137.5052 6.9181 4.5843 3.4425 1.9341 1CL_F 4 0 4 6.3948 1.4077 0.7039 1.9816 4.3052 6.9487 7.3764 3.0712 22.0134 6.2576 6.1006 1.8870 1.8338 0CLD2_F 4 0 4 0.0520 0.0517 0.0259 0.0027 0.0022 0.0421 0.1218 0.1196 99.3637 0.0254 0.0078 0.0620 -3.6724 1LT 4 0 4 0.0599 0.1197 0.0598 0.0143 0.0000 0.0001 0.2395 0.2395 199.6773 0.0004 0.0000 0.0002 -7.7893 4

 A-11
 Table A-xii. Female Rats, Comparisons of One- and Two-Compartment Models

 Individual Animal Results within Dose Groups

 one compartment two compartment

0.1 mg/kg female

Correlation 

(obs vs pred) AIC

Correlation 

(obs vs pred) AIC

0.94 -15.90436

0.955 -18.71725

0.9463 -13.88521

0.1 mg/kg female (extended time course)

0.9748 -22.51555

0.9946 -32.64225

0.9965 -30.0737

0.9914 -27.13124

1.0 mg/kg female intravenous

0.9896 -2.59243

0.9847 6.47658

0.9748 5.75399

0.9538 14.09414

1.0 mg/kg female oral

0.9654 13.33264 0.9952 -1.90618

0.9826 4.95944 0.9828 8.82158

0.9952 3.2396 0.9981 -3.38119

0.9949 -7.62189 0.9949 -3.64701

5 mg/kg female oral

0.9868 40.17303 0.9846 41.67803

0.9846 37.67916 0.9815 40.29375

0.9816 36.31119 0.9947 30.01372

0.9946 26.33912

25 mg/kg female oral

0.9348 123.12264 0.9639 121.62356

0.9641 93.547 0.9641 99.5497

0.9468 99.72258 0.9467 105.71465

0.9873 76.97422 0.9873 82.97019

 A-12
Table A-xiii. Male Rats, One-Compartment Model

Variable N Nmiss Nobs Mean SD SE Variance Min Median Max Range CV%

Geometric Mean

Harmonic Mean

Pseudo SD Mean Log SD 

0.1 mg/kg maleV_F 4 0 4 52.9274 10.5147 5.2573 110.5579 44.1028 50.6338 66.3393 22.2364 19.8662 52.1710 51.4522 9.5789 3.9545 K01 4 0 4 5.5522 6.6896 3.3448 44.7506 0.4407 3.5082 14.7519 14.3112 120.4849 2.3561 1.0479 1.5954 0.8570 K10 4 0 4 0.0041 0.0008 0.0004 0.0000 0.0032 0.0041 0.0050 0.0019 18.7383 0.0040 0.0040 0.0008 -5.5153 BW 4 216.3 11.2beta (L/kg) 0.2450.1 mg/kg male (extended time course)V_F 4 0 4 37.4251 4.3777 2.1888 19.1639 34.0091 36.1936 43.3040 9.2948 11.6971 37.2396 37.0614 4.0574 3.6174 K01 4 0 4 9.1214 4.4205 2.2103 19.5409 3.0705 10.2760 12.8631 9.7927 48.4629 7.9849 6.6324 7.4220 2.0776 K10 4 0 4 0.0042 0.0009 0.0004 0.0000 0.0035 0.0040 0.0053 0.0018 20.7981 0.0041 0.0041 0.0008 -5.4902 BW 4 272.5 5.8beta (L/kg) 0.1371.0 mg/kg male intravenousV 4 0 4 32.5522 1.9014 0.9507 3.6153 29.9308 32.8975 34.4829 4.5521 5.8411 32.5096 32.4661 1.9867 3.4815 K10 4 0 4 0.0106 0.0015 0.0008 0.0000 0.0085 0.0109 0.0121 0.0036 14.2809 0.0105 0.0104 0.0017 -4.5533 BW 4 248.3 9.3beta (L/kg) 0.1311.0 mg/kg male oralV_F 4 0 4 36.9875 5.0695 2.5348 25.6999 33.5521 34.9621 44.4738 10.9217 13.7060 36.7458 36.5234 4.3194 3.6040 K01 4 0 4 1.7823 1.7370 0.8685 3.0173 0.6855 1.0358 4.3723 3.6868 97.4575 1.3350 1.0983 0.6166 0.2890 K10 4 0 4 0.0066 0.0025 0.0013 0.0000 0.0050 0.0055 0.0103 0.0054 38.1633 0.0063 0.0061 0.0016 -5.0667 BW 4 248.7 1.6beta (L/kg) 0.1495 mg/kg male oralV_F 4 0 4 32.3112 4.8117 2.4058 23.1522 26.2113 32.5260 37.9817 11.7704 14.8916 32.0354 31.7538 5.0832 3.4668 K01 4 0 4 0.6954 0.1078 0.0539 0.0116 0.5956 0.6730 0.8399 0.2444 15.5025 0.6893 0.6836 0.0988 -0.3720 K10 4 0 4 0.0068 0.0017 0.0008 0.0000 0.0049 0.0067 0.0090 0.0040 24.1979 0.0067 0.0065 0.0017 -5.0071 BW 4 218.0 3.4beta (L/kg) 0.148225 mg/kg male oralV_F 4 0 4 40.5204 2.9308 1.4654 8.5897 36.8396 40.8789 43.4842 6.6447 7.2329 40.4397 40.3579 3.0151 3.6998 K01 4 0 4 1.5115 0.6513 0.3256 0.4242 0.6062 1.6718 2.0962 1.4900 43.0894 1.3691 1.1990 1.0492 0.3141 K10 4 0 4 0.0071 0.0020 0.0010 0.0000 0.0043 0.0075 0.0090 0.0047 27.9451 0.0068 0.0066 0.0026 -4.9859 BW 225.0 3.7beta (L/kg) 0.1801

 A-13
Table A-xiv. Male Rats, Two-Compartment Model

Variable N Nmiss Nobs Mean SD SE Variance Min Median Max Range CV%

Geometric Mean

Harmonic Mean

Pseudo SD Mean Log SD Log

CGM

0.1 mg/kg oralV1_F 4 0 4 48.4637 12.2283 6.1141 149.5305 39.3870 44.5696 65.3285 25.9416 25.2318 47.3934 46.4259 10.1673 3.8585 0.2399 V2_F 4 0 4 11.1415 7.6879 3.8439 59.1033 2.7865 10.1815 21.4164 18.6299 69.0023 8.8677 6.6436 9.8784 2.1824 0.8478 K01 4 0 4 2.2464 2.0255 1.0127 4.1025 0.4244 2.2807 3.9999 3.5755 90.1638 1.3974 0.8626 0.9233 0.3346 1.2197 CL_F 4 0 4 0.2009 0.0606 0.0303 0.0037 0.1265 0.2096 0.2581 0.1317 30.1333 0.1935 0.1856 0.0692 -1.6427 0.3267 CLD2_F 4 0 4 0.3049 0.1821 0.0911 0.0332 0.0413 0.3730 0.4324 0.3910 59.7312 0.2224 0.1252 0.4264 -1.5032 1.1290 0.1 mg/kg oral extendedV1_F 4 0 4 30.3408 9.0159 4.5080 81.2872 19.2578 30.3915 41.3225 22.0646 29.7156 29.2785 28.1789 10.1562 3.3769 0.3149 V2_F 4 0 4 19.1722 14.7579 7.3790 217.7959 10.0028 12.7323 41.2217 31.2189 76.9754 16.0769 14.2145 6.0913 2.7774 0.6382 K01 4 0 4 2.8750 1.1408 0.5704 1.3015 1.2007 3.2665 3.7664 2.5657 39.6809 2.6356 2.3383 1.9543 0.9691 0.5284 CL_F 4 0 4 0.1397 0.0497 0.0249 0.0025 0.1064 0.1202 0.2120 0.1056 35.5854 0.1340 0.1293 0.0349 -2.0099 0.3224 CLD2_F 4 0 4 0.6509 0.9006 0.4503 0.8111 0.1371 0.2337 1.9993 1.8622 138.3527 0.3474 0.2411 0.1712 -1.0573 1.1997 1 mg/kg intravenousV1 4 0 4 32.7068 2.4987 1.2493 6.2433 30.3884 32.3988 35.6412 5.2527 7.6396 32.6357 32.5654 2.4510 3.4854 0.0760 V2 4 0 4 64.5419 53.4475 26.7238 2856.6356 34.5031 39.5514 144.5620 110.0588 82.8105 52.7869 46.1275 19.3541 3.9663 0.6770 CL 4 0 4 0.1751 0.0417 0.0208 0.0017 0.1170 0.1854 0.2128 0.0959 23.7936 0.1709 0.1662 0.0527 -1.7667 0.2648 CLD2 4 0 4 0.3502 0.1161 0.0581 0.0135 0.2216 0.3454 0.4883 0.2668 33.1568 0.3354 0.3208 0.1180 -1.0925 0.3436 1 mg/kg oralV1_F 4 0 4 25.5409 9.9245 4.9622 98.4956 13.2942 26.9683 34.9329 21.6387 38.8573 23.8838 22.1208 12.0959 3.1732 0.4412 V2_F 4 0 4 17.3278 10.0791 5.0395 101.5874 4.4310 19.0114 26.8573 22.4263 58.1671 14.1875 10.6781 16.3905 2.6524 0.8207 K01 4 0 4 0.9458 1.1459 0.5729 1.3130 0.3389 0.3901 2.6642 2.3253 121.1510 0.6088 0.4732 0.2306 -0.4963 0.9864 CL_F 4 0 4 0.2136 0.0561 0.0280 0.0031 0.1612 0.2004 0.2922 0.1310 26.2483 0.2085 0.2038 0.0476 -1.5680 0.2499 CLD2_F 4 0 4 1.1540 1.2031 0.6016 1.4475 0.0953 0.8290 2.8628 2.7675 104.2553 0.6469 0.2979 1.1733 -0.4355 1.4271 5 mg/kg oralV1_F 4 0 4 25.8013 6.9297 3.4649 48.0208 16.2165 27.7764 31.4357 15.2192 26.8580 24.9914 24.0855 8.9814 3.2185 0.3037 V2_F 4 0 4 38.1791 14.8007 7.4004 219.0613 27.9870 32.6645 59.4005 31.4135 38.7665 36.3128 34.7743 10.4622 3.5922 0.3544 K01 4 0 4 0.5151 0.1483 0.0742 0.0220 0.3092 0.5526 0.6461 0.3369 28.7941 0.4962 0.4749 0.1976 -0.7007 0.3303 CL_F 4 0 4 0.1594 0.0324 0.0162 0.0010 0.1184 0.1622 0.1946 0.0761 20.3126 0.1568 0.1541 0.0356 -1.8531 0.2125 CLD2_F 4 0 4 0.8826 1.5119 0.7560 2.2860 0.0996 0.1402 3.1503 3.0507 171.3095 0.2802 0.1625 0.1015 -1.2721 1.6211 25 mg/kg oralV1_F 4 0 4 37.8559 1.6269 0.8135 2.6468 35.5592 38.3425 39.1797 3.6204 4.2976 37.8292 37.8019 1.6942 3.6331 0.0437 V2_F 4 0 4 33.0604 23.1410 11.5705 535.5047 11.4257 29.3070 62.2020 50.7763 69.9961 26.8232 21.7307 18.0979 3.2893 0.7691 K01 4 0 4 1.3515 0.5876 0.2938 0.3452 0.4909 1.5868 1.7415 1.2505 43.4743 1.2093 1.0310 1.0644 0.1901 0.6063 CL_F 4 0 4 0.2287 0.0616 0.0308 0.0038 0.1615 0.2362 0.2811 0.1197 26.9140 0.2223 0.2158 0.0626 -1.5038 0.2794 CLD2_F 4 0 4 0.2269 0.1410 0.0705 0.0199 0.0985 0.1933 0.4225 0.3240 62.1400 0.1968 0.1721 0.1117 -1.6256 0.6157 

 A-14
 Table A-xv. Male Rats, Comparisons of One- and Two-Compartment Models

 Individual Animal Results within Dose Groups

 one compartment two compartment

0.1 mg/kg male

Correlation

(obs vs pred) AIC

Correlation

(obs vs pred) AIC

0.9157 -64.4 0.9471 -58.6

0.9605 -83.4 0.9605 -79.4

0.9272 -53.6 0.9356 -51.2

0.9497 -64.2 0.9595 -65.4

0.1 mg/kg male (extended time course)

0.8177 17.3 0.905 -8.1

0.9829 -95.7 0.9903 -125.6

0.9804 -96.8 0.9824 -97.9

0.9757 -77.8 0.985 -91.55

1.0 mg/kg male intravenous

0.79 126.7 0.9367 77.9

0.956 81.6 0.9696 72.4

0.9114 91.1 0.914 94

0.9338 89.7 0.9528 82.7

1.0 mg/kg male oral

0.9267 47.7 0.9711 36.2

0.9266 74.5 0.9649 61.5

0.9162 67.2 0.9252 69

0.9413 76.5 0.9366 82.4

5 mg/kg male oral

0.9576 136 0.9587 138.2

0.9399 128.4 0.9659 120

0.8966 156.6 0.9043 158.6

0.9668 139.4 0.9708 139.7

25 mg/kg male oral

0.973 186 0.9822 180

0.9815 187.6 0.9857 183

0.9876 176 0.9877 178.8

0.9732 184.2 0.9822 178.2

 A-15