DEQ-CFW_00001012

Critical Reviews in Toxicology, 34(4):351-384, 2004
Copyright © Taylor and Francis Inc.
ISSN: 1040-8371
DOI: 10.1080/ 10408440490464705
The Toxicology of Perfluorooctanoate
Gerald L. Kennedy, Jr.,'* John L. Butenhoff,2 Geary W. Olsen,2
John C. O'Connor,' Andrew M. Seacat,2 Roger G. Perkins,2
Lisa B. Biegel,3 Sandra R. Murphy,4 and David G. Farrar5
'DuPont, Newark, Delaware, USA; 23M, St. Paul, Minnesota, USA; 3Covance,
Madison, Wisconsin, USA; 4Atofina, Philadelphia, Pennsylvania, USA; and 51neos
Chlor, Manchester, United Kingdom
e Taylor&Francis
healthsciences
Address correspondence to Gerald L. Kennedy, Jr., DuPont Haskell Laboratory, P.O. Box 50, Elkton Road,
Newark, DE, USA. E-mail: [e-mail]
ABSTRACT: PFOA is a peroxisome proliferator (PPAR agonist) and exerts morphological and
biochemical effects characteristic of PPAR agonists. These effects include increased P-oxidation
of fatty acids, increases in several cytochrome P-450 (CYP450)-mediated reactions, and inhibition
of the secretion of very low -density lipoproteins and cholesterol from the liver. These effects
on lipid metabolism and transport result in a reduction of cholesterol and triglycerides in serum
and an accumulation of lipids in the liver. The triad of tumors observed (liver, Leydig cell, and
pancreatic acinar-cell) is typical of many PPAR agonists and is believed to involve nongenotoxic
mechanisms. The hepatocellular tumors observed in rats are likely to have been the result of the
activation of the peroxisome proliferator activated receptor a (PPARa). The tumors observed in
the testis (Leydig-cell) have been hypothesized to be associated with an increased level of serum
estradiol in concert with testicular growth factors. The mechanism responsible for the acinar-
cell tumors of the pancreas in rats remains the subject of active investigation. The mechanism
resulting in the hepatocellular tumors in rats (PPARot activation) is not likely to be relevant to
humans. Similarly, the proposed mechanism for Leydig-cell tumor formation is of questionable
relevance to humans. Acinar tumors of the pancreas are rare in humans, and the relevance of the
these tumors, as found in rats, to humans is uncertain. Epidemiological investigations and medical
surveillance of occupationally exposed workers have not found consistent associations between
PFOA exposure and adverse health effects.
KEYWORDS: Perfluorooctanoate, PFOA, Peroxisome proliferation, Liver tumor, Leydig cell
tumor, Pancreatic acinar cell tumor
Perfluorooctanoate (PFOA, CF3(CF2)6COO-, CAS
No. 3825-26-1) is a perfluorinated carboxylate. As a
commercial product, PFOA has been used primarily
as a surface-active agent in the production of var-
ious fluoropolymers, including tetrafluoroethylene.
Taves et al. (1976) suggested that the organic fluo-
rine that had previously been found in human blood
(Guy, 1973; Taves, 1968a,I968b) was likely PFOA.
Subsequently, trace amounts of organofluorine be-
lieved to be PFOA were detected in the blood of pro-
duction workers and, although the concentrations
were generally in the low parts per million (ppm)
range, they were found to remain in individuals well
after cessation of occupational exposure (Belisle
and Hagen, 1980; Ubel et al., 1980). Improved an-
alytical capabilities using mass spectroscopy were
developed in the 1990s and have allowed the de-
tection of PFOA in serum at part per billion (ppb)
levels (Hansen et al., 2001). Serum samples from
children, adults, and the elderly in the United States
have been shown to contain ppb concentrations of
PFOA (a mean of approximately 5 ppb) (Olsen et al.,
2003a, 2004a, 2004b), and PFOA has been mea-
sured in serum (at concentrations similar to those
observed by Hansen et al., 2001) and in 2/30 liver
samples from organ donors (Olsen et al., 2003b),
351
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suggesting some form of widespread exposure to the
general population. As a result, much attention has
been directed toward evaluating the potential health
consequences associated with the long-term pres-
ence of PFOA in humans, and a risk assessment has
recently been published (Butenhoff et al., 2004a).
This review covers the potential health hazards of
PFOA and does not address potential risk.
Studies conducted with PFOA have ranged
from short to long term, have encompassed a wide
variety of toxicological end points, have involved
several routes of exposure (oral, inhalation, and der-
mal), and have investigated effects in many species
ranging from protozoa to nonhuman primates. In ad-
dition, the potential adverse health effects of PFOA
have been studied in exposed workers through med-
ical surveillance and epidemiological investigation.
This review provides a summary of the current state
of knowledge regarding the potential health hazards
of PFOA and includes pharmacokinetics, toxicity,
and the possible mechanisms through which PFOA
exerts its effects. Current assessments of human epi-
demiology are also provided. The environmental
fate and effects of PFOA are not included in this
review.
PFOA has been used in industry primarily as the
ammonium salt (APFO). In the presence of water,
APFO dissociates readily to PFOA and the ammo-
nium ion. Although most of the major toxicological
studies and industrial exposures involve APFO, the
toxicological effects are likely related to the dis-
sociated anionic form of the acid (PFOA). PFOA
is a member of a broader class of perfluorinated
acids that includes shorter and longer chain perflu-
orinated carboxylates as well as shorter and longer
chain perfluorinated sulfonates. Recently, consider-
able attention has been given to both PFOA and a
sulfonate analog, perfluorooctanesulfonate (PFOS).
The physical properties and the toxicological pro-
file of PFOA and PFOS are different and should
not be confused. This review confines itself to
PFOA.
There are two primary processes used to man-
ufacture PFOA. These are electrochemical fluorina-
tion (ECF) and telomerization. The two processes
differ in that the electrochemical process yields a
product that can contain approximately 15-25%
branched chain PFOA, whereas, the telomer process
yields a linear PFOA. The processes also differ in
the distribution of other chain length compounds as
manufacturing residuals. In addition to being pro-
duced as a commercial product for industrial use,
PFOA also occurs as a contaminant in other ECF or
telomer products.
352
Most of the available, published toxicity infor-
mation for PFOA and its ammonium salt is actu-
ally based on studies using the commercial product
made from the ECF process, FC-143 FLUORAD
brand fluorochemical surfactant. The published lit-
erature frequently refers to this material as APFO or
C-8. FC-143 contained a mixture of the ammonium
salts of several perfluorinated acids as manufactur-
ing residuals. The following is a typical composition
profile:
Ammonium
perfluorooctanoate
(3825-26-1),
(93-97%).
Ammonium
perfluoropentanoate
(68259-11-0),
(1-3 %).
Ammonium
perfluoroheptanoate
(6130-43-4),
(1-3%).
Ammonium
perfluorohexanoate
(21615-47-4),
(1-3 %).
PHARMACOKINETICS
Based on the results of both in vivo and in vitro
studies in rats (Davis et al., 1991; Goecke et al.,
1992; Kuslikis et al., 1992; Ophaug and Singer,
1980; Vanden Heuvel et al., 1991), there is no evi-
dence that PFOA is metabolized in mammals. Con-
sequently, the observed patterns of tissue distribu-
tion and biopersistence in certain species reflect the
net result of absorption and excretion.
Absorption and Biopersistence
PFOA was rapidly absorbed following a sin-
gle gavage administration (25 mg/kg) to rats, and
peak blood levels were attained 1-2 h after dosing
(DuPont, 1982a). Gibson and Johnson (1979) found
that 93% of an oral dose given to male rats was ab-
sorbed within 24 h. There was a clear sex difference
in clearance. Blood levels in female rats showed
>95% clearance 24 h after dosing, while blood lev-
els in males remained relatively high throughout this
period (Table 1). The sex difference in clearance was
even more marked 1 wk after treatment, when blood
levels in males remained relatively high and those
in females had declined to very low levels. Vanden
Heuvel et al. (1991) reported PFOA half-lives of
15 days and < 1 day in male and female rats, respec-
tively, following ip administration. The half-lives
for male and female rats following iv administra-
tion were 5.6 and 0.08 days, respectively (Ohmori
et al., 2003). Importantly, PFOA does not appear
to accumulate in blood of female rats, since the
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TABLE 1
Blood Levels of PFOA in Rats Following Sin-
gle or Multiple Oral Administrations
Blood levels of PFOA (ppm)
Time following
Single-
Single-
25-mg/kg gavage
dose
dose
Dose I V
dose (h)
male
female
female
0.5
23
16
25
2
NA
39
15
8
63
26
13
24
50
0.7
0.8
168
23
0.045
0.1
Note. Data from DuPont (1982a). NA, not available.
"Following 10 previous doses of 25 mg/kg PFOA.
blood profile of an oral dose of 25 mg/kg following
10 previous similar doses was quite similar to that
observed after a single oral dose (Table 1). Fol-
lowing administration of single oral doses of 2.5-
150 mg/kg PFOA to female rats, blood levels at
30 min (3-162 ppm) and 24 h (0.12-18 ppm) were
directly related to the amount of PFOA administered
(DuPont, 1981 a). In rats given either 0, 3, 10, or
30 mg/kg for 28 consecutive days, mean PFOA con-
centrations in the plasma were significantly higher
in male rats. Saturation of binding sites at 30 mg/kg
was suggested (Hanhijarvi et al., 1987).
Blood levels following inhalation exposure of
rats to PFOA exhibited qualitatively similar pat-
terns to those observed after oral exposure (Kennedy
et al., 1986). Thus, following a single 6-h inhala-
tion exposure to PFOA (10 mg/m3), blood profiles
clearly showed that females had a much greater
rate of clearance than males (Table 2). Clearance
in females appeared to be more complete following
PFOA exposures of 0.1 and 1.0 mg/m3, suggest -
ing that the clearance mechanism was overloaded at
10 mg/m3.
The dermal toxicity of PFOA and the pres-
ence of organic fluorine in the blood following der-
mal administration clearly demonstrate the ability
of PFOA to penetrate the skin (Kennedy, 1985).
Excretion and Tissue Distribution
As discussed earlier, the marked sex difference
that exists with regard to clearance is an interesting
and important characteristic of PFOA pharmacoki-
netics in rats. This was shown in a study in which
14C-labeled PFOA was administered iv to male and
female rats (Johnson and Olsen,1980). Females ex-
creted essentially 100% of the administered dose
within the first 24 h after dosing. In contrast, the
males excreted only about 20% of the administered
dose by 24 h and even after 36 days still retained
2.8% and 1.1 % of the total administered 14C in liver
and plasma, respectively, and had detectable levels
in other tissues. This sex difference in elimination
by male and female rats is reported by other inves-
tigators (Davis et al., 1991; Hanhijarvi et al., 1982;
Kojo et al., 1986; Kudo et al., 2001; Ophaug and
Singer, 1980; Vanden Heuvel et al., 1991; Ylinen
et al., 1989). Approximately 25% of the radiolabel
in females and 10% in males appears in the feces
and could represent either unabsorbed PFOA or that
involved in biliary excretion.
An additional comparative study with male and
female mice, hamsters, and rabbits dosed by gavage
with 10 mg/kg 14C-PFOA and sacrificed after 5 days
(hamsters, mice) or 1 wk (rabbits) showed that sex
differences in the rate of clearance of PFOA are not
restricted to rats (DuPont, 1982a). In the case of
hamsters, the sex difference was exactly opposite
to that observed in rats. The male hamster excreted
about 99% of the administered PFOA in 5 days while
TABLE 2
Blood Levels of PFOA in Rats Following Inhalation Exposure
Time (h) after 6-h
inhalation exposure to
concentration indicated
Blood levels of PFOA (ppm)
Females
0.1 mg/m3 1.0 mg/m3 10.0 mg/m3
Males,
10.0 mg/m3
0.5
2
7
109
137
2
2
17
69
157
8
0.85
4
71
182
24
0.14
0.56
52
147
Note. Data from Kennedy et al. (1986).
353
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TABLE 3
Distribution of Radiolabel Following Gavage Administration of 14C-PFOA to Both Sexes of
Rats, Mice, Hamsters, and Rabbits
Percent of radiolabel administered'
Rat
Mouse Hamster
Rabbit
Matrix
Male
Female
Male Female Male Female
Male
Female
Urine
25.6
73.9
3.4 6.7 90.3 45.3
76.8
87.9
Feces
9.2
27.8
8.3 5.4 8.2 9.3
4.2
4.6
Tissue
59.6
0.6
73.6 50.0 0.7 26.5
<0.1
0.3
Note. Data from DuPont (1982a).
"Each animal received a single gavage dose of 10 mg/kg 14C-PFOA. Rats, mice, and hamsters were sacrificed at 5 days
and rabbits 7 days after administration.
the female hamster excreted 58% of the adminis-
tered dose in the same time period. In rabbits, both
sexes excreted the PFOA as rapidly and completely
as the female rat and male hamster, while both sexes
of mice behaved more like the male rat, retaining
50-70% of the administered radiolabel 5 days af-
ter treatment (Table 3). In dogs, Hanhijarvi (1988)
found plasma half-lives of 473 and 541 h in 2 male
dogs and 202 and 305 h in 2 female dogs following a
single iv injection. After a single iv dose of PFOA,
male and female cynomolgus monkeys showed a
possible difference in elimination rate, with elimi-
nation half-lives being somewhat longer in females
(mean = 32.6 ± 8.0 days) than in males (mean =
20.9 f 12.5 days) (Noker, 2003).
Excretion occurs in both the urine and the fe-
ces, the main difference between fast and slow ex-
cretors being the rate of urinary excretion. Thus,
urinary excretion in the female rat, male hamster,
and both sexes of the rabbit accounts for 74-90%
of the total administered radiolabel compared with
that in mice (3-7%), male rats (26%), and female
hamsters (45%). Less than 5% of the radiolabel was
found in expired air. Biliary excretion was slower in
male than female rats (Kudo et al., 2001), and less
than 1 % was reported excreted in bile of either male
or female rats (Vanden Heuvel et al., 1991).
As indicated in Table 3, combined tissue
residues at the time of sacrifice are inversely related
to the rate of urinary excretion and represent <0.7%
of the administered dose with female rats, male ham-
sters, and rabbits (both sexes). Measurements of the
distributions of the 14C-label in the slow -excreting
species (Table 4) indicated that blood, liver, and kid-
ney were the tissues containing the largest residues.
Lower concentrations were found in lung, heart,
skin, and testis, and only traces were found in mus-
cle, fat, and brain.
354
In a study in which PFOA (0, 3, 10, and
30 mg/kg) was administered to rats by gavage for
28 days, major deposition sites were serum, liver,
and kidney (Ylinen et al., 1990). These tissues were
also identified in rats as the major deposition sites
by Davis et al. (1991) following oral administration.
In the female, deposition at the 3-mg/kg dose level
was greater in the liver than the kidney. This was
reversed at higher doses, suggesting the existence
of a saturable renal excretory mechanism.
Pregnant and nonpregnant rats showed similar
blood levels of PFOA following single oral or in-
halation exposures (DuPont, 1982a), and there was
no evidence to suggest that PFOA accumulated in
the blood following repeated exposures. PFOA was
shown to cross the placenta to the fetus (Table 5),
and measurable residues were found in fetal spleen,
heart, lungs, and fat following a single gavage dose
of 10 mg/kg 14C-PFOA on day 19 of pregnancy
(DuPont, 1982b).
Tissue Binding
Tissue distribution of PFOA is undoubtedly dic-
tated to some extent by its ability to bind avidly
to plasma and other proteins. As early as 1954,
Klevens and Ellenbogen provided data to support
PFOA-induced protein -protein aggregation through
a proposed mechanism of van der Waals complex-
ing between polar groups and the fluorocarbon chain
(Klevens and Ellenbogen, 1954). Nordby and Luck
(1956) reported the unusually high ability of PFOA
to reversibly precipitate protein under conditions
of mild acidity through albumin binding to hydro-
gen ions and perfluorooctanoate ions. Binding and
precipitation with albumin were maximal (better
than trichloroacetic acid) at a pH of about 4, where
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TABLE 4
Tissue Distribution of Radiolabel in Slow -Excreting Species Following Gavage Administra-
tion of 14C-PFOA in Rats, Mice, Hamsters
Tissue
Male
rat
µg Equivalents of PFOA per g (ml) wet weight'
Mouse
Male Female
Female
hamster
Blood
23.5
13.8 10.1
8.8
Liver
40.0
43.2 45.3
7.3
Kidneys
24.0
2.91 2.2b
7.1
Lungs
8.7
1.4b 1.3b
3.8
Heart
6.4
1.2b 0.6b
2.9
Skin
4.8
3.5 0.2
3.4
Testes
3.2
0.9b -
-
Muscle
1.9
1.1 0.5
0.9
Fat
1.7
1.6 1.3
1.5
Brain
0.6
0.2b 0.8b
0.3
Note. Data from DuPont (1982a).
"Each animal received a single gavage dose of 10 mg/kg 14C-PFOA and was sacrificed 5 days after administration
Calculations based on specific activity of 14C-PFOA.
bRepresents µg equivalents for entire organ.
PFOA, a strong acid (pKa 1-2; Moroi et al., 2001)
is mainly (>99%) in its anionic form and where
there are positive charges on albumin. The bind-
ing was reportedly reversible. In contrast, Vanden
Heuvel et al. (1992a) reported covalent binding of
14C-PFOA to proteins in liver, plasma, and testes of
rats. Covalent binding, however, was defined sim-
ply by the fact that some of the 14C label adminis-
tered to rats in vivo or added to in vitro preparations
was not extractable into organic solvents. Current
evidence suggests that greater than 90% of PFOA
would be bound to blood albumin in the rat, monkey,
and human (Han et al., 2003; Kerstner-Wood et al.,
TABLE 5
Transplacental Transfer of PFOA
2003). Reduction of the in vitro binding of PFOA
to albumin in the presence of cysteine (Peterson
et al., 1991; Vanden Heuvel et al., 1992a) suggests
the possible involvement of sulfhydryl (-SH) groups
on the protein, although other thiol-containing com-
pounds did not affect binding and there was no ev-
idence of the formation of PFOA conjugates with
glutathione. Similarly, Kuslikis et al. (1992) did not
find evidence of covalent binding of PFOA to coen-
zyme A thiol. Ylinen et al. (1990) found 99% of
PFOA to be bound to serum protein in rats. Plasma
protein binding was >98% in rats (Ohmori et al.,
2003).
µg Equiv. of PFOA
per g (ml) Tissue'
Time
Maternal
following
Maternal
Fetal
blood/
dose (h)
blood
tissue
fetal ratio
2
14.5
0.7
22
4
22.4
7.5
7.7
8
12.5
4.5
4.5
Note. Data from DuPont (1982b).
'Each animal received a single gavage dose of 10 mg/kg 14C-PFOA at day 19 of pregnancy. Calculations based on specific
activity of 14C-PFOA.
355
DEQ-CFW 00001016

Mechanism of Elimination
The biochemical mechanism(s) for the ob-
served sex and species differences in excretion of
PFOA remain to be elucidated. Since PFOA is not
metabolized, it seems probable that the mechanism
involves differences in clearance through the kid-
ney. Hanhijarvi et al. (1982) and Vanden Heuvel
et al. (1991) suggested that the rapid urinary ex-
cretion of PFOA in female rats involves a specific
renal active transport mechanism and there is some
evidence for the involvement of steroid hormones.
Vanden Heuvel et al. (1992b) showed that testos-
terone inhibited renal excretion of PFOA in male but
not female rats. Ylinen et al. (1989) had previously
reported that estradiol increased urinary excretion
in both castrated and intact male rats. The marked
species differences complicate any simple hormone -
mediated explanation. A report by Kojo et al. (1986)
in which male and female weanling rats were dosed
by gavage for 28 days demonstrated a sex differ-
ence in serum concentrations of PFOA in that male
serum concentrations were 6-19 times higher than
those in females. However, sex differences in re-
nal clearance, as measured on days 7 and 28, were
only statistically significant in the low dose group
(3 mg/kg) on day 7, and females appeared to be as
affected, if not more responsive, than males. Assum-
ing that these weanling rats had not reached sexual
maturity, these latter observations suggest that the
sex differences in renal clearance observed in adult
rats is a result of sexual maturation.
The difference in urinary excretion of PFOA
between male and female rats may be due to al-
tered expression by the organic anion transporter
proteins OAT2 and OAT3 (Kudo et al., 2002). Kudo
et al. (2002) have shown that clearance of PFOA
can be decreased in both sexes by treatment with
probenecid, by testosterone treatment of castrated
males, or by treatment of females with testosterone.
Increased clearance of PFOA followed castration
or treatment of males with estradiol. Curiously,
ovariectomy increased renal clearance, which could
be reduced by either estradiol or testosterone treat-
ment. The authors of this study demonstrated that
renal clearance of PFOA in rats is inversely propor-
tional to peroxisomal beta -oxidation, suggesting a
link to peroxisome proliferation. They also demon-
strated that the differential expression of OAT2 and
OATS transporters in liver and kidney of male and
female rats might be under hormonal control.
Goecke-Flora and Reo (1996) examined the ef-
fect of carbon chain length on the excretion of a ho-
mologous series (C7—C11) of perfluorinated acids
356
in rats. They showed that the C7 and C8 acids were
excreted primarily in the urine, the C10 and C11
derivatives mainly in the bile and the C9 acid in
both urine and bile, mainly the latter. Johnson et al.
(1984) reported a 9.8-fold increase in fecal elimi-
nation when cholestyramine (4% in diet) was fed to
male rats for 14 days after administration of a single
iv dose of 14C-PFOA (13.3 mg/kg). Levels of 14C in
liver, plasma, and red blood cells of cholestyramine-
treated rats 14 days after administration of PFOA
were about 50% of those in controls, which sug-
gests that biliary excretion and enterohepatic recir-
culation can be modified.
Human Biopersistence
Contrary to the relatively rapid rates of elimina-
tion reported in laboratory animals (DuPont, 1982a;
Johnson and Ober, 1980; Butenhoff et al., 2002),
PFOA appears to be slowly eliminated in humans
(Ubel et al., 1980; Burris et al., 2002). In an in-
terim report, Burris et al. (2002) reported the serum
half-life of elimination for PFOA was 4.4 years (SD
3.5) in 9 retired fluorochemical production employ-
ees who had serum concentrations measured over
an 18-mo time period. Initial concentrations ranged
from 0.1 to 1.8 ppm. Although nonoccupational
sources of PFOA exposure during the follow-up
time period were not accounted for and the half-
life study is still in progress, the slow elimination
rate observed suggests that humans have the least
ability to eliminate PFOA of any species studied.
ACUTE TOXICITY STUDIES
Acute Oral Toxicity
The results of several acute oral LDS() stud-
ies conducted with PFOA are relatively consistent
(Table 6) and indicate the material to be of moderate
acute oral toxicity. The data reported in Table 6 are
from studies conducted with at least five animals per
sex per dose and at least three dose levels. Data from
several studies conducted to determine minimum
acute lethal dose are not included but are consistent
with the studies presented in Table 6. With the ex-
ception of the high male LD50 observed in one study
in which water was the vehicle, the toxicity of PFOA
is similar when administered in either corn oil or wa-
ter. Despite the much longer biological half-life of
PFOA in male versus female rats, the LD50 is ap-
proximately the same in each sex (430-680 mg/kg)
and is also similar to that observed in mice. While
DEQ-CFW 00001017

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TABLE 6
Acute Toxicity of PFOA
Number/
LD50 (mg/kg),
Species
sex/dose
Vehicle
*LC50(mg/L)
Reference
Oral
Rat
5 (M/F)
Acetone (40%)
680 (M), 430 (F)
Griffith and Long (1980)
Corn oil (60%)
Rat
10 (M/F)
Corn oil
470 (M), 482 (F)
DuPont (1981d)
Rat
10 (M)
Corn oil
478
DuPont (1981c)
Rat
5 (M/F)
Water
1800 (M), 600 (F)
Hazleton (1997)
Mouse
10 (M/F)
Corn oil
457
DuPont (1981e)
Guinea pig
10 (M/F)
Corn oil
178 (M), 217 (F)
DuPont (1981f)
Dermal
Rat
5 (M/F)
Aqueous paste
7000 (M), >7500 (F)
DuPont (1979a),
Kennedy(1985)
Rat
5 (F)
Aqueous paste
>7500
DuPont (1979a)
Rabbit
5 (M)
Aqueous paste
4300
DuPont (1979a),
Kennedy(1985)
Rabbit
5 (M/F)
Aqueous paste
>2000
Hazleton (1995c)
Inhalation
Rat
6 (M)
4-h exposure
0.98*
Kennedy et al. (1986)
Rat
5 (M/F)
1-h exposure
No mortality 18.6*
Griffith and Long (1980)

there are no differences in the sensitivity of castrated
or ovariectomized versus intact rats (male or female)
to PFOA (DuPont, 1981b), newborn rats (<2 days
old) (LD50 ^250 mg/kg) appear to be more sen-
sitive to PFOA than weanlings and adult animals
(LD50 340-580 mg/kg) (DuPont, 1983a). In the rat,
acute oral exposure to higher doses generally re-
sults in enlarged livers, elevations of serum enzyme
levels, gastrointestinal irritation, and weight loss.
Guinea pigs appear to be somewhat more suscep-
tible than other rodents with an acute oral LD50 of
178-217 mg/kg. A study with three male dogs also
indicated lethality 48 h following a single oral dose
of 450 mg/kg PFOA (DuPont, 1965).
Several other studies have been conducted to
determine the effects of PFOA alone and in combi-
nation (pre- and postexposure) with other chemicals
or drugs. Pretreatment of rats with phenobarbital
(enzyme inducer) or proadifen hydrochloride (Ca2+
channel blocker) did not change the LD50 of PFOA
(DuPont, 1981c), while, as previously discussed, a
single treatment with cholestyramine reduced mor-
tality compared to that in rats dosed with PFOA
alone (DuPont, 1982c).
Acute Dermal Toxicity
As indicated by the data in Table 6, the acute
dermal toxicity of PFOA applied as an aqueous paste
is relatively low in both rats and rabbits, with LD50
values of 7000 mg/kg (or greater) and 4300 mg/kg,
respectively (DuPont, 1979a). Death typically oc-
curred 3-7 days after exposure. Clinical signs of tox-
icity included body weight loss, wet and/or stained
perineal area, lethargy, labored breathing, diarrhea,
and severe skin irritation, with necrosis at the higher
dose levels (DuPont, 1979a).
Acute Inhalation Toxicity
PFOA was moderately toxic when administered
to rats by inhalation (head only) as a dust. The
approximate 4-h lethal concentration in rats was
0.81 mg/L and the LC50 was 0.98 mg/L (Table 6)
(Kennedy et al., 1986); all deaths occurred within
48 h of exposure. Exposures (4 h) to concentra-
tions between 0.38 and 0.83 mg/L PFOA caused
weight loss and increased liver -to -body weight ra-
tio, which, in surviving animals, returned to normal
by the end of a 42-day recovery period. During ex-
posure, rats displayed irregular breathing and red
discharge around the eyes and nose. Additionally, all
rats exposed to >0.81 mg/L PFOA showed corneal
358
opacity and corrosion, and dying rats had hyperin-
flated lungs.
SKIN AND EYE IRRITATION
AND SKIN SENSITIZATION
Dermal Irritation
No irritation was observed in a study (Griffith
and Long,1980) in which PFOA powder (0.5 g) was
applied to dry and moistened abraded skin sites on
rabbits and left occluded for 24 h. Subsequent stud-
ies, however, have established that PFOA (0.5 g)
applied to the shaved skin of rabbits as an aqueous
paste for 24 h caused moderate erythema and mild
edema that persisted for about 48 h (Kennedy, 1985;
Hazleton, 1990). These data suggest that PFOA
should be classified as a mild skin irritant. Rabbits
are more sensitive to the dermal irritation effects of
PFOA than rats.
Dermal Sensitization
A dermal sensitization test (Buehler method)
conducted with guinea pigs has determined that
PFOA is not a contact sensitizer (Moore, 2001).
Eye Irritation
PFOA produces moderate eye irritation char-
acterized by corneal, iridal, and conjunctival ef-
fects. Instillation of solid PFOA Q8.5-100 mg) into
rabbit eyes caused moderate corneal opacity, iri-
tis, and conjunctivitis (DuPont, 1979b; Griffith and
Long, 1980). The effects gradually receded over
time, and prompt washing of the eye reduced the
effects and provided a more rapid recovery. In ad-
dition to the eye irritation studies in which PFOA
was instilled directly into the eye, rats exposed to
PFOA particulate (0.81 mg/L) during a 4-h inhala-
tion period (head only) exhibited corneal opacity
and ulceration, which was microscopically evident
42 days postexposure (Kennedy et al., 1986).
SUBCHRONIC TOXICITY STUDIES
Oral Exposure
A summary of the results of the most impor-
tant subchronic studies conducted with PFOA in
mice, rats and monkeys is shown in Table 7. The
DEQ-CFW 00001019

w
co
TABLE 7
Subchronic Toxicity Studies with PFOA
Species/study type
Number/sex/dose
Dose level, range of dietary
concentration (ppm) or mg/kg
Results (mg/kg)
Reference
Oral route
Mouse
5 (M/F)
10-10,000 ppm (1.5-1500 mg/kg)"
Mortality (100%) at >450 mg/kg; mortality
DuPont (1981g)
14-Day feeding
at 150 mg/kg; increased liver weight/
body weight ratio at >4.5 mg/kg
Mouse
5 (M/F)
0.01-3000 ppm (0.0015—
Mortality (100%) at >450 mg/kg; mortality
Kennedy (1987)
14/21-Day feeding
450 mg/kg)"
and weight loss at >45 mg/kg; increased
liver weight/body weight ratio at
>4.5 mg/kg
Mouse
5 (M/F)
30-30,000 ppm (1.5-1500 mg/kg)"
Mortality (100%) at > 150 mg/kg;
Griffith and Long (1980)
28-Day feeding
weight loss (males > 5 mg/kg; females
> 1.5 mg/kg); increased liver weight
in males and females at > 1.5 mg/kg
Rat
5 (M/F)
30-30,000 ppm (1.5-1500 mg/kg)°
Mortality (100%) at >500 mg/kg; decreased
Griffith and Long (1980)
28-Day feeding
body weight (males >50 mg/kg; females
> 150 mg/kg); increased liver weight/
body weight ratio (males > 1.5; females
>50 mg/kg)
Rat
5 (M/F)
10-1000 ppm (0.5-50 mg/kg)"
Weight loss and increased liver weight
Griffith and Long (1980)
90-Day feeding
at > 15 mg/kg; panlobular diffuse
hepatocellular hypertrophy at >50 mg/kg
Rat
55 (M)
1-100 ppm (0.05-5 mg/kg)'
Weight loss at 5 mg/kg; increased liver
Perkins (1992)
90-Day feeding and
weight, and hepatocellular hypertrophy
8-wk recovery
at >0.5 mg/kg and palmitoyl CoA oxidase
at > 1.5 mg/kg—reversible during
recovery period; NOEL = 0.05 mg/kg
Monkey (rhesus)
2 (M/F)
3-100 mg/kg
Mortality (100% at 2-5 wk) at 100 mg/kg;
Griffith and Long (1980)
90-Day gavage
body weight decrease and mortality
(1 M and 2 F during wk 7-12) at
30 mg/kg; atrophy of lymphoid tissues
at >30 mg/kg; NOEL = 3 mg/kg
(Continued on neat page)

w
rn
0
TABLE 7
Subchronic Toxicity Studies with PFOA (Continued)
Dose level, range of dietary
Species/study type Number/sex/dose concentration (ppm) or mg/kg Results (mg/kg) Reference
Monkey (cynomolgus)
6-Month gavage and
90-day recovery
Inhalation route
Rat
10-Day inhalation and
28- to 84-day recovery
Dermal Route
Rat
10 Applications
dermal and 84-day
recovery
Rabbit
10 Applications
dermal and 14-day
recovery
4 or 6 (M)
24 (M)
15 (M)
10 (M/F)
°Approximate dose equivalent to dietary concentration.
0.3-20/30 mg/kg (Dosage lowered Liver weight increases at all dose levels;
from 30 to 20 mg/kg due to poor body weight loss at 20/30 mg/kg
tolerance
1, 8, 84 mg/m3 (6 h/day)
20-2000 mg/kg (6 h/day for 2 wk,
5 days/wk)
100 mg/kg (6 h/day for 2 wk,
5 days/wk)
Mortality (2) at 84 mg/m3 and weight loss
that recovered by day 16; increase in liver
weight, serum alkaline phosphatase,
hepatocellular hypertrophy, and necrosis
observed at 84 and 8 mg/m3 but all
recovered by day 42; focal or multifocal
hepatocellular necrosis seen in some
animals at high and mid dose;
NOAEL = 1 mg/m3
Skin irritation and reversible reduction in
body weight at >200 mg/kg (persistent
body weight loss at 2000 mg/kg);
increased liver weight, increase in AST
and ALT, hepatocellular hypertrophy and
necrosis at >20 mg/kg
Reversible reduction in body weight but
elevated blood fluorine levels throughout
recovery period
Butenhoff et al. (2002)
Kennedy et al. (1986)
Kennedy (1985)
Riker (1981)

results of the various studies are quite consistent
and show that mortality occurs in mice and rats at
dietary levels equivalent to about 150-450 mg/kg
(DuPont, 1981 g; Kennedy, 1987; Griffith and Long,
1980). Nine repeat doses (gavage) caused mortality
in mice at about 10 mg/kg (DuPont, 1983b). The
primary target organ is the liver as is shown by in-
creased liver weight, increased serum transaminase
enzyme activity (ALT, AST), and diffuse hepato-
cellular hypertrophy accompanied at high doses by
acidophilic degeneration and/or necrosis. In rats, ef-
fects in males are typically more pronounced than
in females, a fact that may reflect, in part, the more
rapid elimination of PFOA by females.
In a 90-day feeding study with male rats fed ei-
ther 0, 1, 10, 30, or 100 ppm (approximately equiv-
alent to 0, 0.05, 0.5, 1.5, and 5 mg/kg), liver weight
changes were the most sensitive indicator of re-
sponse, along with hepatocellular hypertrophy oc-
curring at 0.5 mg/kg or higher. Body weight gain
decreases were seen at 1.5 mg/kg or higher, and the
no -observed -effect level (NOEL) for all effects was
0.05 mg/kg (Perkins, 1992).
In a study with rhesus monkeys (Griffith
and Long, 1980), mortality was observed at both
100 mg/kg (2-5 wk) and 30 mg/kg (7-12 wk). All
monkeys in these dose groups were anorexic and had
emesis, black stool, pale face and gums, swollen face
and eyes, hypoactivity, and prostration. Histopathol-
ogy included moderate atrophy of lymphoid folli-
cles in the spleen and lymph nodes. As a result of
these signs and symptoms, Griffith and Long (1980)
suggested that the major sites of activity in the rhe-
sus monkey were in the gastrointestinal tract and
reticuloendothelial system rather than liver as ob-
served in rodents and in a 6-mo oral study in male
cynomolgus monkeys that is discussed in the fol-
lowing paragraph.
In another study, male cynomolgus monkeys
(n = 4 or 6) received daily oral (capsule) doses
(0, 3, 10, and 30/20 mg/kg) of PFOA for 26 wk
(Butenhoff et al., 2002). The initial high dose of
30 mg/kg was poorly tolerated and treatment was
discontinued within 2 wk. Reinitiating of treatment
at 20 mg/kg continued to result in serious weight
and appetite loss in all but 2 of the 6 monkeys in this
group. One 30/20-mg/kg dose group monkey died,
showing hepatocellular degeneration and necrosis,
and 3 others were taken off treatment prior to sched-
uled termination. Additionally, 1 monkey in the
3-mg/kg group died near the end of the 6-mo treat-
ment period. Despite intensive investigation, cause
of death remains uncertain, as does the relationship
to PFOA treatment. Dose -dependent increases in
liver weight occurred in all PFOA-treated groups.
The lowest -observed -effect level (LOEL) based on
liver weight increase was 3 mg/kg. At 6 mo, there
was no increase in peroxisome proliferation relative
to controls in any of the treated monkeys sacrificed
at this time. A thorough battery of clinical chemistry,
biochemical, and hormonal tests was conducted and
appeared unaffected. There were no PFOA-related
macroscopic or microscopic changes, and no effects
on clinical chemistry, urinanalysis, or hematology
were seen. With the exception of the two monkeys
that did not survive to term, there were no PFOA-
related macroscopic or microscopic changes, and no
effects on clinical chemistry, urinalysis, or hematol-
ogy were seen. Liver weight increase may have been
due, in part, to increased mitochondrial prolifera-
tion. No increased cell proliferation was observed
in the liver, pancreas, and testes.
In repeated exposure studies (in some cases also
in acute studies) a sex difference was demonstrated
with PFOA being a potent peroxisome proliferator
in male rats, less potent in female rats (Sohlenius
et al., 1992). This sex difference was not seen in
mice. The induction in male rats was strongly depen-
dent on testosterone levels (Kawashima et al.,1989).
These effects progressed to liver hepatomegaly, par-
ticularly in male rats (Takagi et al., 1992), and were
shown to be modified by coadministration of estro-
gen (Cameron et al., 1982).
PFOA was fed for 7 days to wild -type and
PPARcx-null (knockout) mice and measures of im-
munocompetence were conducted (Yang et al.,
2002a). Reductions in spleen weight and spleno-
cyte number along with reductions in thymus weight
and thymocyte number were observed in the wild -
type mice. Splenic changes were absent and thymic
changes were less pronounced in the PPARa-
null mice. This response was also seen with the
peroxisome proliferating agents, WY-14,643, di-
(2-ethylhexyl)phthalate, and nafenopin, suggesting
that PPARu activation may play a role in the im-
munomodulation caused by peroxisome prolifera-
tors (Yang et al., 2000, 2002a, 2002b). The changes
in the spleen and thymus were rapidly reversible
when the splenic and thymic responses returned to
normal by 5 and 10 days postdosing, respectively
(Yang et al., 2001).
Male C57131/6 mice fed 200 ppm PFOA for up
to 10 days were immunized with horse red blood
cells, and the Immoral response was measured by
the plaque -forming cell assay and determination of
the antibody titer by enzyme -linked immunosorbent
assay (ELISA). PFOA treatment prevented both the
increase in plaque -forming cells (anti-IgM and IgG)
361
DEQ-CFW 00001022

and increase in serum levels of IgM and IgG nor-
mally increased by such immunization (Yang et al.,
2002b). Ex vivo spleen cell proliferation in response
to both T- and B-cell activation was attenuated by
PFOA feeding, but the analogous in vitro treat-
ment of mouse spleen cells had no effect, suggest-
ing an indirect mechanism of action (Yang et al.,
2000).
Inhalation Exposure
Subchronic inhalation studies in rats with
PFOA clearly indicate the liver to be the pri-
mary target organ (Table 7). In a 10-day, repeated -
exposure study, increased mortality and weight loss
at 84 mg/m3 was associated with increased liver
weight, increases in serum enzymes indicative of
liver damage, and hepatocellular hypertrophy and
necrosis (Kennedy et al., 1986). In surviving rats,
the liver effects seen at 84 and 8 mg/m3 were
not present by recovery day 42. The no -observed -
adverse -effect level (NOAEL) for all effects was
I mg/m3.
Dermal Exposure
While rats treated dermally for 10 days with
200 and 2000 mg/kg PFOA survived, they showed
a dose -dependent increase in body weight loss
and intermittent red discharge from eyes and nose
(Kennedy, 1985). An increase in liver weight and
serum transaminases, as well as hepatocellular hy-
pertrophy and necrosis, was observed at >20 mg/kg;
lowered kidney and spleen weights were also ob-
served at these doses. With the exception of those
associated with the high dose, all effects had dis-
appeared within a 42-day period following treat-
ment. A reversible reduction in body weight was
also reported following 10 daily dermal appli-
cations of 100 mg/kg PFOA to rabbits (Riker,
1981).
DEVELOPMENTAL TOXICITY
The developmental toxicity of PFOA has been
studied in rats and rabbits by the oral exposure
route and in rats by the inhalation exposure route
(Table 8) (Gortner, 1981,1982; Staples et al., 1984).
In those studies, pregnant animals were treated with
graded doses/exposures of PFOA during organogen-
esis. The potential developmental effects of PFOA
on the fetuses were evaluated both externally and
362
internally, and by skeletal examination of the fe-
tuses obtained prior to natural delivery. For one set
of oral and inhalation studies in rats, dams were al-
lowed to deliver and pups were observed through the
lactation period. These studies, as discussed later,
demonstrate that maternal exposure to PFOA during
organogenesis does not result in embryofetal toxic-
ity or developmental abnormalities in the offspring
of rats and rabbits.
The first developmental study to be conducted
with PFOA was a study in rats by Gortner (1981).
In this study, maternal toxicity was observed at the
highest dose (150 mg/kg) and consisted of body -
weight reductions and mortality Q of 22 dams). Re-
productive organs were unaffected by treatment. Ex-
amination of fetuses did not reveal any increase in
embryofetal toxicity or structural abnormalities that
were attributable to PFOA treatment. Lens abnor-
malities, originally attributed to PFOA treatment,
were found subsequently to be an artifact of the sec-
tioning technique (Staples, 1985).
In another oral study, rats were given 100 mg
PFOA/kg of body weight by gavage from gestation
day 6 through 15 (Staples et al., 1984). One group
of 25 pregnant rats and their litters were examined
on day 21 of gestation. Another group of 12 treated
dams gave birth and the resulting pups were exam-
ined on day 35 postpartum. Maternal effects includ-
ing death and decreased maternal body weight gain
were seen in both groups. No developmental toxi-
city or abnormalities were seen in the fetuses, and
offspring showed normal viability and growth.
By the inhalation route, groups of pregnant rats
were exposed to concentrations of either 0.14, 1.2,
9.9, or 21 mg PFOA/m3, 6 h/day from day 6 through
15 of gestation (Staples et al., 1984). Exposure to
the 21 mg/m3 resulted in the death of 3 of 12 rats,
with the suviving dams showing toxicity as evi-
denced by reduced weight gains and clinical signs
including lethargy and chromodacyorrhea. Reduced
weight gains were also seen in dams exposed to
9.9 mg/m3. No effects were seen in those exposed
to either 0.14 or 1.2 mg/m3. While mean fetal body
weights were reduced in fetuses of surviving dams
exposed to either 9.9 or 21 mg/m3, most likely this
was the result of maternal toxicity and is not con-
sidered primary embryofetal toxicity. There were
no structural abnormalities in fetuses from any of
the exposure groups that could be associated with
PFOA exposure.
In a rabbit developmental study (Gortner,
1982), rabbits were given oral doses of either 1.5,
5, or 50 mg PFOA/kg from gestation day 6 through
18. The number of rabbits producing litters in this
DEQ-CFW 00001023

TABLE 8
Summary of Studies on Developmental/Reproductive Toxicity of PFOA
Study type
Species
Route/dosing
days
Dose/exposure
levels (mg/kg)
Number of
litters
Findings
Developmental'
Rat
Oral
0
20
Maternal —none
GD 6-15
Fetal —none
0.05
21
Maternal —none
Fetal —none
1.5
19
Maternal —none
Fetal —none
5
21
Maternal -none
Fetal —none
150
14
Maternal —mortality
and body weight
gain reductions
Fetal —none
Developmentalb
Rat
Oral
0
25
Maternal —none
GD 6-15
Fetal —none
100
25
Maternal —mortality
and body weight
gain reductions
Fetal —none
Developmentalb
Rat
Inhalation
0
24
Maternal —none
GD 6-15
Fetal —none
0.14 (mg/m3)
24
Maternal —none
Fetal —none
1.2 (mg/m3)
24
Maternal —none
Fetal —none
9.9 (mg/m3)
15
Maternal —body weight
gain reductions
Fetal —none
21 (mg/m3)
12
Maternal —mortality
and body weight
gain reductions
Fetal —body weight
reductions
Developmental`
Rabbit
Oral GD
0
18
Maternal —none
6-18
Fetal —none
1.5
17
Maternal —none
Fetal —none
5
18
Maternal —none
Fetal —none
50
18
Maternal —body
weight gain reductions
Fetal —none
Reproductive
Rat
Oral
0
30160e
Fetal —none
(Premating—
Parental —none
2nd generation)
Offspring —none
1
30/60e
Parental—fertilty
normal; liver effects
in males
Offspring —none
(Continued on next page)
363
DEQ-CFW 00001024

TABLE 8
Summary of Studies on Developmental/Reproductive Toxicity of PFOA (Continued)
Route/dosing Dose/exposure Number of
Study type Species days levels (mg/kg) litters Findings
3 30/60e Parental —fertility
normal, liver,
body weight effect
in males
Offspring —none
10 30/60e Parental —fertility
normal, liver,
body weight effects
in males
Offspring —none
30 30/60e Parental —fertility
normal, liver,
body weight effects
in males, minimal
kidney weight effect
in females
Offspring —reduced
viability pre and
post weaning 1st
generation; not 2nd
(preweaning only
examined); pup
weight reduction,
delay in vaginal
opening, preputial
separation
Note. No structural anomalies associated with PFOA dosing/exposure.
"From Gortner (1981).
dFrom Staples et al. (1984).
`From Gortner (1982).
'From Butenhoff et al. (2004b).
e30 Per sex per group P generation; 60 per sex per group Fl generation.
study was low in all groups, a fact that compro-
mises interpretation of the study. A reduction in
maternal body weight gain was observed in rabbits
given 50 but not 1.5 or 5 mg/kg. No other signs
of response to PFOA were observed in the preg-
nant rabbits. The number of fetuses per litter from
all treatment groups was as expected, fetuses were
structurally normal, and weighed essentially the
same as their untreated counterparts. No evidence
was seen of either embryotoxicity or teratogenicity.
A dose -dependent increase was noted in the number
of fetuses with the natural and stress -related varia-
tion of 13th ribs. This latter finding is known to be
quite variable, is not a malformation per se, and is
believed to be a transient effect that remodels post-
364
natally (Christian et al., 1987). In addition, it is be-
lieved that this variation is not likely to be relevant to
humans.
REPRODUCTIVE TOXICITY
A two -generation reproduction study in rats
was conducted with PFOA (Butenhoff et al., 2004).
Rats were treated with oral gavage doses of either 1,
3, 10, or 30 mg/kg body weight/day, In the parental
rats, signs of toxicity were observed at all dose lev-
els in the males and at 30 mg/kg in females. In
males, body weight gain suppression was observed
DEQ-CFW 00001025

at all doses (except 1 mg/kg in the P, generation),
along with organ weight changes (liver, kidney, and
spleen). Female parental rats were relatively unaf-
fected by treatment, with decreased kidney weights
seen in P, females and decreased weight gains in
FI females only at 30 mg/kg. There were no effects
on any of the mating or fertility parameters in ei-
ther generation. At 30 mg/kg, a number of effects
in the offspring were observed, including decreased
pup weights, increased pup mortality (Fl generation
only), and delayed vaginal opening and preputial
separation. These findings were not observed at any
of the lower doses. Clearly, reproductive success
(i.e., mating and fertility) was not compromised in
rats at dosages of up to 30 mg/kg.
With respect to offspring, at the highest dose
tested (30 mg/kg), decreased pup weights during
lactation and increased pup mortality were observed
in the F, but not the FZ generation. Preweaning mor-
tality was increased numerically in FI pups but was
not statistically significant. Postweaning mortality
was statistically significant in the Fl pups but was
not evaluated in the F2 generation offspring (all F2
offspring were necropsied at weaning).
The increased incidence of pup mortality at
30 mg/kg in the Fl generation is most likely a result
of a general failure of the offspring to thrive, based
on the apparently compromised nutritional status
of the offspring at pre- and/or postweaning as re-
flected by reduced body weight. Eleven of the 13 Fl
offspring that died postweaning died prior to post -
weaning day 8, and these included the 9 lightest
pups, suggesting that low body weight was a fac-
tor. Although not statistically significant at all time
points, mean pup body weights were consistently
decreased throughout the lactation period (90, 90,
89, 92, and 95% of control on postnatal days 1, 5, 8,
15, and 22, respectively). Body weight effects in off-
spring have also been observed in reproduction stud-
ies performed with other peroxisome proliferators
such as gemfibrozil, RMI 14,514, and hydrochlo-
rofluorocarbon 123 (HCFC-123) (Fitzgerald et al.,
1987; Gibson et al., 1981; Malinvemo et al., 1996).
It seems likely that the compromised nutritional sta-
tus of some offspring resulting in low body weight
is responsible for the increased pup mortality ob-
served in the two -generation reproduction study
with PFOA.
The data from this study shows delayed age at
preputial separation in males (mean = 3.7 days) and
delayed age at vaginal opening in females (mean =
1.7 days) in the Fl offspring. The delays in sex-
ual maturation may have been the result of delayed
growth of the Fj offspring. As noted earlier, pup
weights were consistently decreased throughout the
lactation period.
While the body weights of the Fl generation
offspring were similar to the controls at the time
of sexual maturation, it is plausible that the delayed
growth that was observed early in lactation may have
contributed to the delays that were observed in sex-
ual maturation of the F, offspring. Decreased body
weights can result in nonspecific delays in puberty
(Carney et al., 1998; Glass et al., 1976; Glass and
Swerdloff, 1980; Kennedy and Mitra, 1963; Marty
et al., 1999, 2001 a, 2001b, 2001c; Ronnekleiv et al.,
1978; Stoker et al., 2000a, 2000b; Widdowson and
McCance, 1960). In a recent report by Lewis et al.
(2002), variability of sexual maturation data was
evaluated in control populations of Sprague-Dawley
rats. They found that the typical variability among
control groups was approximately 2 days, a finding
that was also consistent with the typical variability in
age at sexual maturation reported by others (Ashby
and Lefevre, 2000; Clark, 1999; Marty et al., 1999;
Stoker et al., 2000b). Since nonspecific effects on
body weight can cause general delays in sexual mat-
uration, interpreting delays in sexual maturation can
be problematic in studies where generalized delays
in growth occur, such as those that were observed
in the current study of PFOA. It is also possible
that the delay in preputial separation may have been
the result of elevated estradiol and decreased testos-
terone; however, these effects have only been shown
in adult male rats. As noted previously, PFOA did
not compromise reproductive success (i.e., mating
and fertility) in rats at dosages of up to 30 mgfkg.
Notably, the overall results of the first and second
generation appear to be similar in that there was no
decrease in reproductive success in the Fj generation
or increase in adverse effects in the FZ generation.
CHRONIC TOXICITY AND
ONCOGENICITY
The oncogenicity of PFOA has been investi-
gated in two separate 2-yr feeding studies in Sprague
Dawley rats. In the first study (Riker, 1987), PFOA
was administered to male and female rats at dietary
levels of 0, 30, and 300 ppm. In the second study
(Biegel et al., 2001), PFOA was administered to
male rats at a dietary level of 300 ppm, and the
study included both ad libitum and pair -fed con-
trol groups. PFOA was found to increase the in-
cidence of benign hepatocellular, testicular Leydig-
cell, and pancreatic acinar-cell tumors at 300 ppm in
a study by Biegel et al. (2001), and PFOA increased
365
DEQ-CFW 00001026

TABLE 9
Summary of Hyperplasia/Neoplasia in Liver, Testes, and Pancreas of Rats Fed PFOA
Organ/lesion
Control (fed ad libitum),
incidence (%)
Pair -fed control,
incidence (%)
PFOA (300 ppm),
incidence (%)
Liver
Adenoma (A)
2/80 (3)
1/79 (1)
10/76 (13)b
Carcinoma (C)
0/80 (0)
2/79 (3)
0/76 (0)
A + C
2/80 (3)
3/79 (4)
10/76 (13)b
Testes
Leydig-cell hyperplasia
11/80 (14)
26/78 (33)
35/76 (46)a
Leydig-cell adenoma
0/80 (0)
2/78 (3)
8/76 (11)b
Pancreas
Acinar-cell hyperplasia
14/80 (18)
8/79 (10)
30/76 (39)b
Acinar-cell adenoma (A)
0/80 (0)
1/79 (1)
7/76 (9)b
Acinar-cell carcinoma (C)
0/80 (0)
0/79 (0)
1/76 (1)
A + C
0/80 (0)
1/79 (1)
8/76 (11)b
Note. Data from Biegel et al. (2001).
"Significantly different from ad libitum control.
bSignificantly different from pair -fed control.
testicular Leydig cell tumors at a dietary dose of 300
ppm in the study by Riker Pharmaceuticals (Riker,
1987). The incidence of these tumors is presented
in Table 9. In the following discussion, each tumor
type is discussed in turn, and the potential mecha-
nisms are covered in a later section of this report.
Hepatocellular Adenoma
In the study by Biegel et al. (2001), histopatho-
logical evaluation revealed PFOA-related increases
in hepatocellular adenoma. Hepatocellular adenoma
occurred at an incidence of 13% (10/76), as com-
pared to 3% (2/80) and i% (1/79) in ad libitum and
pair -fed controls, respectively. A similar finding was
not observed in the Riker study (Riker, 1987).
Testicular Leydig-Cell Tumors
Both studies showed increases in hyperplasia
and benign tumors (adenoma) of testicular Leydig
cells. In the Riker (1987) study, the incidence of
Leydig-cell adenomas was 0/50, 3/50, and 7/50 at
dosages of 0, 30, and 300 ppm PFOA, respectively.
Biegel et al. (2001) found an increase in the in-
cidence of Leydig-cell hyperplasia and adenomas,
with adenoma incidences of 0/80, 2/78, and 8/76
in the ad libitum controls, pair -fed controls, and
300 ppm PFOA group, respectively.
366
Pancreatic Acinar-Cell Tumors
In the Biegel et al. (2001) study, an in-
crease was observed in pancreatic acinar-cell ade-
noma and combined pancreatic acinar-cell ade-
noma/carcinoma. Acinar-cell adenoma incidence
was 0/80, 1/79, and 7/76 for the ad libitum con-
trols, pair -fed controls, and 300 ppm dose group,
respectively. One acinar-cell carcinoma was present
in the 300 ppm dose group (1/76), yielding an over-
all adenoma/carcinoma incidence of 8/76. The Riker
(1987) study did not result in an increase in pancre-
atic tumors, although a recent (2001) peer review
of pancreatic tissues from both studies revealed ev-
idence of acinar-cell hyperplasia in the Riker study
but not an increase in adenoma (S. R. Frame, per-
sonal communication).
Mammary Tumors
In the Riker (1987) study, the incidence of fi-
broadenomas of the mammary gland apparently was
increased in female rats (22%, 42%, and 48% at
0, 30, and 300 ppm in diet, respectively). There was
no apparent difference in incidence despite dietary
levels that were an order of magnitude apart. The
authors of this study concluded that the mammary
tumor data did not reflect an effect of PFOA. Riker
Pharmaceuticals did not have an adequate histori-
cal control database for comparison. However, un-
treated control rats (same strain and supplier) from
DEQ-CFW 00001027

13 chronic toxicity/oncogenicity studies conducted
at Haskell Laboratory (DuPont) from 1984 to 1987
provided 947 control rats. Historical control data
for this period is also available from the supplier,
Charles River.
The difference in the incidence of fibroadeno-
mas in the PFOA-treated groups versus the Haskell
Laboratory historical controls is not statistically sig-
nificant (p = .3). The incidence of fibroadenomas in
the 13 referenced Haskell Laboratory studies ranged
from 24 to 54% with a mean of 37%. In the Riker
study, the control group incidence was just below
and the test group incidences were near the top of the
study control range. The incidences in the PFOA-
treated groups (42 and 48%) were similar to the
average of the Haskell Laboratory historical control
groups (37%). Further, historical control data posted
on the Charles River Laboratories web site report an
average fibroadenoma incidence among 24 studies
of 41 % with a range of 13-61 %. These data further
support the Riker study authors' conclusion that the
distribution of fibroadenomas in the PFOA study
were a reflection of background incidence and were
not related to PFOA treatment.
PFOA was a positive modulator of hepa-
tocarcinogenesis in rats initiated by diethylni-
trosamine or diethylnitrosamine followed by 2-
acetylaminofluorene treatment (Abdellatif et al.,
1991). In an initiation —selection —promotion assay
in rats, PFOA was found to act as a promoter of
liver tumors (Nilsson eta]., 1991).
GENETIC TOXICITY
The weight of evidence from studies evaluat-
ing the genotoxicity of PFOA indicates that PFOA
is not genotoxic. These studies include evaluation
of mutagenicity, clastogenicity, and cell transforma-
tion (Table 10).
PFOA has not shown a potential to induce
DNA point mutations or recombinations. PFOA
has shown a lack of activity in bacterial reverse
mutation assays including Salmonella typhimurium
and Escherechia coli strains and in yeast (Saccha-
romyces cerevisiae) recombination assays in both
the absence and the presence of metabolic activation
(Litton Bionetics, 1978; Hazleton, 1995a, 1996a).
Similarly, in the Chinese hamster ovary (CHO) for-
ward mutation assay, PFOA did not induce a statis-
tically significant increase in the number of mutant
colonies in the treated cells (Toxicon, 2002).
Chromosomal aberrations were assessed in hu-
man lymphocytes and CHO cells. PFOA did not in-
duce significant increases in the numbers of chromo-
somal aberrations in human lymphocytes (Hazleton,
1996b; Notox, 2002). When tested in CHO cells,
PFOA was classified as being clastogenic only for
highly toxic treatment levels of sufficiently short du-
ration in which the cells did not disintegrate prior
to the scheduled harvest times (Hazleton, 1996c,
1996d). In view of the cytotoxicity observed, the
biological significance of this positive response is
questionable.
PFOA did not induce a significant increase in
bone marrow polychromatic erythrocytes after oral
administration to mice (Hazleton, 1995b, 1996e).
There was no evidence of cell transformation us-
ing the C311 10Tl/2 cell line observed at any of
the tested dose levels (Stone Research Laboratories,
1981).
HUMAN EXPOSURE AND
EPIDEMIOLOGY
The presence of organic, covalently bound fluo-
rine in human blood was reported by Taves (1968a,
1968b) and Taves et al. (1976) tentatively identi-
fied a component of the organic fluorine as PFOA.
Following that finding, 3M began monitoring its
fluorochemical production workers (Ubel et al.,
1980). Workplace monitoring was originally based
on determination of total organofluorine in serum.
Speciation and quantitation of PFOA rather than
total organofluorine was introduced in 3M's med-
ical monitoring program in the 1990s. 3M has
also surveyed sera from the general (nonoccupa-
tional) populations. Using high-performance liquid
chromatography—electrospray tandem mass spec-
trometry, Hansen et al. (2001) and Olsen et al.
(2003a, 2004a, 2004b) have shown in three U.S.
general population studies serum concentrations of
PFOA, regardless of age, averaging approximately
5 ppb with an upper bound of the 95th percent tol-
erance limit approximating 10 to 15 ppb. In another
study, Olsen et al. (2003b) examined a total of 30 hu-
man donor livers for the presence of PFOA. Almost
all donor livers were below the lower limit of quanti-
tation, which ranged between 5.4 and 35.9 ng/g. Of
the 2 donor livers measured above the lower limit
of quantitation, 1 liver had a PFOA concentration
of 2.5 ng/g and the other had a mean analysis of
46.9 ng/g.
The first published report of serum concentra-
tions in an ammonium perfluorooctanoate (APFO)
production workforce (Ubel et al., 1980) indicated
a total organic fluorine content of 1 to 71 ppm in
367
DEQ-CFW 00001028

TABLE 10
Summary of Studies on Genotoxicity of PFOA
Study type
Study description Results Reference
In vitro mutagenicity,
Assayed with S. typhimurium Negative Litton (1978)
Ames test
(TA 1535, TA 1537, TA 1538,
and TA 100) and S. cerevisiae
D4 yeast with and without
metabolic activation
In vitro mutagenicity,
Assayed with S. typhimurium Negative Hazleton (1995a)
Ames test
(TA 1535, TA 1537, TA 98,
and TA 100) and E. coli
(WP2uvrA) with and without
metabolic activation
In vitro mutagenicity,
Assayed with S. typhimurium
Negative
Hazleton (1996a)
Ames test
(TA 1535, TA 1537, TA 98,
and TA 100) and E. coli
(WP2uvrA) with and without
metabolic activation
In vivo mutagenicity,
5 Mice/sex dosed with 1250,
Negative
Hazleton (1995b)
mouse micronucleus
2750, and 5000 mg/kg and
bone marrow evaluated after
24, 48, and 74 h
In vivo mutagenicity,
5 Mice/sex dosed with 200,
Negative
Hazleton (1996e)
mouse micronucleus
400, and 800 mg/kg and bone
marrow evaluated after 24,
48, and 72 h
Chromosomal aberration
Assayed in Chinese hamster
Negative in two
Hazleton (1996c),
ovary (CHO) cells with and
replicates and
Hazleton (1996d),
without metabolic activation,
weakly positive
Toxicon (2002)
in one
Chromosomal aberration
Assayed in human
Negative
Hazleton (1996b)
lymphocytes in whole blood
Notox (2002)
with and without metabolic
activation
Cell transformation
Assayed with mammalian C3H
Negative
Stone research
10T1/2 cells
Laboratories
(1981)
3M employees at an electrochemical fluorination
plant in Cottage Grove, MN. The highest serum
concentrations were observed in employees with the
longest production work history and among a subset
of workers whose jobs involved drying and pack-
aging operations. Using gas chromatographic tech-
niques, Ubel et al. reported that 90% of the organic
fluorine was PFOA. Additional analyses of serum
total organic fluorine or PFOA measurements have
been conducted since Ubel et al. in conjunction with
medical surveillance of these Cottage Grove produc-
368
tion workers (Gilliland,1992; Gilliland and Mandel,
1996; Olsen et al., 1998,2000,2003c). Serum PFOA
concentrations during the 1990s were comparable
to those initially estimated by Ubel et al. (1980)
with means (range in parentheses) of 5.0 ppm (0-80
ppm), 6.8 ppm (0.0-114), and 6.4 ppm (0.1-81 ppm)
in 1993, 1995, and 1997, respectively (Olsen et al.,
2000).
Medical surveillance of these Cottage Grove
production workers has included a medical ques-
tionnaire; measurement of height and weight;
DEQ-CFW 00001029

pulmonary function testing; and standard biochemi-
cal and hematology tests (Ubel et al.,1980; Gilliland
et al., 1996; Olsen et al., 2000, 2003c). Hep-
atic and lipid -related blood tests have included
alanine aminotransferase (ALT), aspartate amino-
transferase (AST), alkaline phosphatase, gamma-
glutamyl transferase (GGT), total and direct biliru-
bin, total cholesterol, high -density lipoproteins
(HDL), low -density lipoproteins (LDL), and triglyc-
erides. Ubel et al. (1980) initially reported that no
health problems were associated with exposure to
fluorochemicals in this workforce, although aggre-
gate analyses were not presented. Others have subse-
quently not shown consistent abnormal associations
between clinical chemistry or hematology findings
with either total organic fluorine levels (Gilliland
and Mandel, 1996) or serum PFOA concentrations
(Olsen et al., 2000, 2003c) in this workforce. Al-
though Gilliland and Mandel (1996) suggested that
serum total organic fluorine levels might modulate
hepatic responses to obesity and alcohol, these in-
teractions were not observed in three subsequent
surveillance examinations that analyzed specifically
for serum PFOA concentrations (Olsen et al., 2000).
In response to toxicological findings that sug-
gested PFOA might modulate endocrine activ-
ity in the rat (Biegel et al., 1995, 2001), serum
levels of several hormones (estradiol, free and
bound testosterone, follicle -stimulating hormone
[FSH], luteinizing hormone [LH], prolactin, and
thyroid -stimulating hormone [TSH]) were incor-
porated in the medical surveillance program of
these Cottage Grove workers starting in 1990.
Other hormones (dehydroepiandrosterone sulfate,
17-hydroxyprogesterone, and sex -hormone -binding
globulin) were added in 1993 and 1995. Gilliland
(1992) reported a positive nonlinear association
with estradiol and a negative nonlinear association
with free or bound testosterone in relation to serum
total organic fluorine concentrations. These associ-
ations were not confirmed with analyses specific to
serum PFOA concentrations, although mean estra-
diol levels were 10% greater among employees with
the highest serum PFOA concentrations (30 ppm)
(Olsen et al., 1998). Possible explanations for the
disparity between findings may involve the use of
different biomonitoring and hormonal assay meth-
ods and possible misclassification of confounding
variables, most importantly body mass index (Olsen
et al., 1998). Based on their data, Olsen et al. (1998)
concluded there was reasonable assurance of no sub-
stantial hormonal changes associated with PFOA at
the serum levels measured among these male pro-
duction employees. Limitations of these hormonal
investigations include their cross -sectional design
and the lower levels of serum PFOA measured, com-
pared with those concentrations reported to cause
effects in laboratory animal studies. In 2000, sev-
eral thyroid tests were analyzed including TSH, thy-
roxin (T4), free T4, triiodothyronine (T3), and thy-
roid hormone binding ratio (Olsen et al., 2003b). In
a linear model, total T4 was negatively associated
with serum PFOA concentrations but the variation
explained was minimal (r2 = 0.3). None of the T4
values were below the reference range of the assay.
No associations were observed for free T4, TSH, or
total T3. There were no statistically significant dif-
ferences in the percentage of test results that were
above or below these other thyroid hormone refer-
ence ranges either.
It has been hypothesized that a sustained in-
crease in cholecystokinin (CCK), resulting from
hepatic cholestasis, maybe the mechanism by which
peroxisome proliferators like PFOA cause pancreas
acinar-cell adenomas in rats (Obourn et al., 1997;
Biegel et al., 2001). Even though there has been
no significantly increased risk of pancreatic can-
cer mortality in this workforce (Alexander et al.,
2001), levels of plasma CCK-33 (CCK associated
with 33 amino acids is the predominant form) were
assayed as part of the 1997 medical surveillance pro-
gram (Olsen et al., 2000). Cholecystokinin (mean
28.5 pg/ml, SD 17.1, range 8.8-86.7 pg/ml) ap-
proximated the assay's reference range (up to 80
pg/ml) for a 12-h fast and was negatively, not posi-
tively, associated with employees' serum PFOA lev-
els; thus, the epidemiologic data did not support the
toxicological positive association between CCK and
PFOA.
Serum PFOA concentrations have been re-
ported in other 3M workplace male and female
populations in Antwerp, Belgium and Decatur,
Alabama (Olsen et al., 2003d). The primary focus of
this research endeavor, however, was PFOS as serum
PFOA concentrations (mean 1 to 2 ppm, range
0.0 to 13 ppm) have been historically lower than
those reported for 3M Cottage Grove APFO pro-
duction workers. Several important demographic
(age higher in Decatur) and lifestyle factors (body
mass index higher in Decatur, alcohol consumption
higher in Antwerp) were different between these
two populations. A non -adjusted analysis of serum
PFOS concentrations by male employee quartiles
was associated with higher concentrations of PFOA
(Olsen et al., 2003d). The highest PFOS quartile
(mean 2.7 ppm, range 1.7-10.0 ppm) had a mean
serum PFOA concentration of 2.7 ppm (range 0.3-
12.7 ppm). This highest quartile had significantly
369
DEQ-CFW 00001030

(p < 0.05) higher mean serum triglyceride, alka-
line phosphatase, ALT and T3 levels than the low-
est quartile (mean serum PFOA concentration of
0.54 ppm). These associations were confounded by
the fact that the Decatur population was proportion-
ately more represented in the highest quartile than
the lowest quartile. In a longitudinal analysis over a
6 year time period (1994-2000), serum PFOA con-
centrations were positively associated with choles-
terol as well as triglycerides. This association was
considered to be inconsistent with the hypolipi-
demic effect of this compound in rats that is thought
to be associated with the activation of the PPAR
receptor.
Ubel et al. (1980) and sequentially Gilliland
and Mandel (1993) and Alexander et al. (2002)
have examined the mortality experience of employ-
ees working at the 3M Cottage Grove manufactur-
ing site. The study cohort definition has changed
slightly over time as a minimum 6-mo employ-
ment eligibility was used until Alexander et al.
incorporated a 1-yr eligibility criterion. Gilliland
and Mandel (1993) followed the vital status of
2788 male and 749 female Cottage Grove workers
from 1947 through 1989 and reported no statisti-
cally significant increased cause -specific standard-
ized mortality ratios (SMR). Cumulative exposure
to PFOA was estimated using the surrogate mea-
sure of months of chemical division employment.
This surrogate measure was not specific, however, to
those APFO production areas (e.g., drying and pack-
aging) where exposure was the highest. Prostate
cancer mortality was associated with length of em-
ployment in the chemical division in a proportional
hazard analysis. Ten years of employment was as-
sociated with a 3.3-fold increase (95% Cl 1.0-10.6)
in prostate cancer mortality relative to that of work-
ers not employed in the chemical division. However,
only one employee had worked directly in the PFOA
production building (Olsen et al., 1998).
Alexander et al. (2002) addressed the ma-
jor limitation of the Gilliland and Mandel (1996)
mortality study, the lack of job and department
specificity in the duration of employment analy-
ses, by computerizing all work history records of
Cottage Grove employees with at least 1 yr of cu-
mulative employment. Alexander et al. (2001) then
constructed a calendar -year, job- and department -
specific exposure matrix from this computerized
database. The necessity of incorporating job and de-
partment specificity into an exposure matrix was
demonstrated by Olsen et al. (2003d), who did
not find an association between employees' serum
PFOA concentrations and their years worked at
370
this Cottage Grove manufacturing site but did find
associations with specific production ( job) areas.
Alexander et al. then had a panel of veteran work-
ers and plant industrial hygienists categorize the job
and department titles into three exposure groups:
1. Definite PFOA exposure (potentially high).
These jobs included workers employed in the
areas where electrochemical fluorination, dry-
ing, shipping, and packaging of PFOA occurred
throughout the history of the plant.
2. Probable PFOA exposure. These jobs included
other chemical division jobs where exposure to
PFOA was possible but with lower or transient
exposures.
3. Not exposed to fluorochemicals.
In total, 3183 male and 809 female workers were
followed for vital status from 1947 through 1997.
The all -cause mortality (SMR = 0.9, 95% CI 0.8-
0.9) and all -cancer mortality (SMR = 0.8, 95% Cl
0.7-1.0) ratios for the entire study population re-
gardless of exposure classification, as well as for
the exposure subcohorts, were less than expected
than the general population. Specifically, there was
no association between cohort members with a min-
imum of 1 yr employment in a job with definite or
probable PFOA exposure and all -cancer mortality
(SMR = 0.9, 95% Cl 0.7-1.1), cancer of the liver
(SMR = 0.6, 95% CI 0.0-3.3), pancreas (SMR =
1.4, 95% CI—0.5-3.1) or prostate (SMR = 1.2,
95% CI 0.4-2.5), or cirrhosis of the liver (SMR=
0.7, 95% CI 0.2-1.8). Contrary to the findings from
Gilliland and Mandel (1993), prostate cancer mor-
tality was not associated with duration of employ-
ment among those with definite or possible expo-
sure to PFOA (observed/expected in parentheses):
0—<1 yr (0/0.1), 1—<5 yr (2/1.4), 5—<10 yr (0/0.8),
and > 10 yr (4/2.9). The SMR was 1.4 (95% Cl
0.4-3.5) for prostate cancer in the > 10 yr duration
category. Deaths from cerebrovascular disease ex-
ceeded the number of expected deaths in the defi-
nite PFOA exposed subgroup: 5 observed and 1.9
expected (SMR = 2.6, 95% CI 0.8-6.0). The ba-
sis for this observation remains to be elucidated.
Although the Alexander et al. (2002) investigation
improved upon the methods used for exposure as-
sessment, some misclassification of exposure was
likely. Maintenance and other mobile workers not
specifically identified as definitely PFOA-exposed
workers may have routinely entered the areas of
high exposure (drying and packaging). The extent to
which this misclassification occurred and the effects
on the study results is unknown.
DEQ-CFW 00001031

A retrospective cohort mortality study has also
been performed on the 3M Decatur workforce pop-
ulation (Alexander et al., 2003). The exposure of
interest was PFOS although exposure to PFOA
was also possible. Three bladder cancers occurred
among workers who held a high PFOS exposure
classified job (SMR 12.8, 95% Cl2.6-38.4). It is not
clear whether these three deaths were attributed to
fluorochemical exposure, an unknown bladder car-
ccinogen encountered during the course of work,
and/or non -occupational exposures. This bladder
cancer association was not observed in the Cot-
tage Grove mortality study as there was 1 death
(SMR = 0.7, 95 % CI 0.0-3.8) among those employ-
ees who had a minimum of one year employment
in a job with definite or probable PFOA exposure
(Alexander, 2001).
MECHANISMS OF ACTION
AND HUMAN RELEVANCE
Systemic Toxicity
Despite its general lack of chemical reactivity,
PFOA exerts a wide range of biological activity in
rodents leading to both acute and subchronic tox-
icity and mortality and chronic effects manifested
by increased incidences of tumors in liver, testes
and pancreas. Signs and symptoms of toxicity are
often nonspecific and include weight loss, lethargy
and aphagia, and a general wasting syndrome. There
are, however, marked species and sex differences in
pharmacokinetics and toxicity that complicate gen-
eral conclusions regarding the mechanism of action
and potential human risk. A large number of bio-
chemical effects have been reported, and it is often
difficult to determine whether these reflect a primary
biochemical lesion or are secondary effects result-
ing from interaction with some other target. There
remains a good deal of mechanistic uncertainty at
this time as discussed in the following sections.
The liver is a primary target organ for both
short-term and chronic effects of PFOA in rats
(Griffith and Long, 1980; Olson and Anderssen,
1983; Kennedy, 1985; Pastoor et al., 1987) and
cynomolgus monkeys (Butenhoff et al., 2002). The
increased liver weight in both species does not ap-
pear to be a result of hepatocellular hyperplasia
(no increase in nuclear DNA) and has been vari-
ously attributed to increases in peroxisomes, endo-
plasmic reticulum, and mitochondria (Ikeda et al.,
1985; Pastoor et al., 1987; Butenhoff et al., 2002;
Berthiaume and Wallace, 2002; Biegel et al., 2001).
Higher doses lead to liver degeneration and necrosis
and the appearance in the serum of enzymes reflect-
ing liver damage.
Biochemical Effects
Biochemical changes associated with PFOA
treatment reflect the observed changes in hepatic
cell morphology. Thus, in rats, the increase in
smooth endoplasmic reticulum of the liver is linked
with increases in total cytochrome P-450 (CYP450)
content and benzphetamine N-demethylase activ-
ity, while cytoplasmic enzymes such as glutathione
S-transferase and UDP-glucuronyltransferase were
unaffected (Pastoor et al., 1987). PFOA is also a
potent inducer of hepatic CYP450, CYP450 re-
ductase, and epoxide hydrolase in mice (Permadi
et al., 1992) and enhances hepatic carboxylesterase
activity in rats (Hosokawa and Satoh, 1993). The
ability of perfluoro compounds to bind to the sub-
strate binding sites and activate transcription factors
for CYP450 and to both inhibit and induce several
CYP450 isozymes and their associated detoxica-
tion activities has been recognized for some time,
even though they are not themselves metabolized
(Ullrich and Diehl, 1971; Armstrong and Lowe,
1989). It is not yet clear how much of the total
induction of CYP450 involves the isozymes asso-
ciated with xenobiotic metabolism and how much
involves the specific isozyme, CYNA1, that is spe-
cific for the w-hydroxylation of lauric acid and is
characteristically induced by peroxisome prolifera-
tors (discussed later). Both Obraztsov et al. (1993)
and Gross and Rudiger (1991) have suggested that
the effects must be critically dependent on the sol-
ubility and/or intercalation of the chemicals in the
membrane lipids.
Peroxisome Proliferation and
Effects on Lipid Metabolism
PFOA and other perfluoroalkanoic acids are
part of a widening group of substances including
plasticizers (phthalate esters), hypolipidemic drugs
(clofibrate, nafenopin), and some herbicides (phe-
noxyacetic acids), solvents (trichloroethylene), and
naturally occurring compounds such as long -chain
fatty acids that are known as PPARa agonists (Ikeda
et al., 1985; Just et al., 1989; Pastoor et al., 1987;
Cook et al., 1992, 1994; Biegel et al., 1995, 2001;
Cattley et al., 1998). PFOA has been shown to acti-
vate the PPARa receptor but not the PPARy receptor
371
DEQ-CFW 00001032

(Maloney and Waxman, 1999). As with all PPs,
treatment of rodents with PFOA initiates a char-
acteristic sequence of morphological and biochem-
ical events in the liver and, to a lesser extent, the
kidney. These events include marked hepatocellu-
lar hypertrophy due to an increase in number and
size of peroxisomes, a large increase in peroxiso-
mal fatty acid ,8-oxidation, an increased CYP450-
mediated w-hydroxylation of lauric acid, and vari-
ous changes in lipid metabolism (Ikeda et al., 1985;
Pastoor et a1.,1987; Berthiaume and Wallace, 2002).
This response is initiated by the activation of the
nuclear hormone receptor PPARa (Green, 1995;
Ashby et al., 1994; Lake, 1995).
PPARa is able to increase the transcription rate
of responsive genes and is the major mediator of PP
in rodent liver. The critical role of PPARa in PP in
mice has recently been clearly established. PPARa-
null mice do not show the typical PP -mediated re-
sponses or signs of hepatic hyperplasia or neoplasia
(adenomas or carcinomas) in chronic studies with
PPs (Peters et al., 1997; Ward et al., 1998). Long-
term exposure of rodents to PPs characteristically re-
sults in an increased incidence of liver tumors (Doull
et al., 1999; IARC, 1995). Pronounced species dif-
ferences have been reported following treatment of
animals with PPs in vivo and have been observed
in hepatocyte cultures in vitro (Ashby et al., 1994;
IARC, 1995; Bentley et al., 1993; Elcombe et al.,
1997; Lake, 1995; Maloney and Waxman, 1999).
Rats and mice are highly, perhaps uniquely, respon-
sive to the effects of PPs, whereas Syrian hamsters
exhibit an intermediate response and guinea pigs
seem to be practically nonresponsive, as are pri-
mates, including both Old World and New World
(e.g., marmoset) species, and humans (Bentley et al.,
1993; Tucker and Orton, 1993; Graham et al., 1994;
Kurata et al., 1998; Pugh et al., 2000; Butenhoff
et al., 2002). There are, however, differences in the
effects exerted by different groups of PPs; For exam-
ple, PFOA did not cause hepatocellular hyperplasia
(like most other PPs; Pastoor et al., 1987), although
the study design did not include looking at early time
points, such as 3-21 days. In rats, the hepatocellu-
lar response is primarily hypertrophic and is caused
by an increase in the number of peroxisomes, with
proliferation of the smooth endoplasmic reticulum.
In addition, there is evidence for proliferation of mi-
tochondria that may, in part, account for increased
liver mass (Butenhoff et al., 2002; Berthiaume and
Wallace, 2002).
PFOA interferes with fatty acid metabolism and
cholesterol synthesis in the liver (Haughom and
Spydevold, 1992). Serum cholesterol levels were
372
reduced 50-70% after only 24 h in rats fed diets
containing 0.02% PFOA, and triglycerols were re-
duced to about 60% of controls after 7 days. Mea-
surements of selected enzymes in hepatocytes from
the PFOA-treated rats showed significant decreases
in HMG-CoA reductase, the rate -limiting step in
cholesterol biosynthesis, and in acyl coenzyme A
(CoA) cholesterol acyltransferase (ACAT), the en-
zyme responsible for the esterification of choles-
terol. It was concluded that the hypolipidemic effect
of PFOA is due, in part, to the reduced synthe-
sis and esterification of cholesterol, combined with
the enhanced oxidation of fatty acids in the liver
(Haughom and Spydevold,1992). Kudo et al. (1999)
confirmed the earlier observation by Pastoor et al.
(1987) that PFOA treatment increased the level of
triglycerides in the liver and linked this with in-
creases in two triglyceride-synthesizing enzymes,
glycerol 3-phosphotransferase and diacylglycerol
acyltransferase. It also seems possible, however, that
the increase in triglycerides in the liver as well as
their decrease in the plasma following treatment
with PFOA is associated with a decrease in triglyc-
eride secretion from the liver. Clearly, PFOA has the
ability to disrupt lipid metabolism by any of several
mechanisms that at the present time are not well
understood. The importance of these, if any, in the
toxicity of PFOA is not known.
PFOA, like other peroxisome proliferators, has
been shown to uncouple oxidative phosphorylation
(Keller et al., 1992; Starkov and Wallace, 2002).
While PFOA has been shown to induce mitochon-
dria) proliferation in rats (Berthiaume and Wallace,
2002) and monkeys (Butenhoff et al., 2002), this
finding may not be tied to peroxisome proliferation
(Berthiaume and Wallace, 2002). In isolated mito-
chondria, PFOA is a weak inhibitor of mitochondrial
bioenergetics (Starkov and Wallace, 2002). At a rel-
atively high concentration (100 µM) PFOA exerted
a small stimulatory effect on both resting and state
3 respiration, induced a small decrease in the A* of
mitochondria, and increased the respiratory control
index; it was without effect on the ADP:O ratio. It is
probable that at high concentrations, PFOA acts as a
strong detergent to disrupt the inner membrane and
decrease respiratory control without uncoupling ox-
idative phosphorylation. The effects on mitochon-
dria may relate to oxygen stress as well as apoptosis
(Keller et al., 1992; Panaretakis et al., 2001; Permadi
et al., 1993).
In the recent subchronic study with cynomol-
gus monkeys, one high -dose animal (30 mg/kg) de-
veloped severe symptoms of aphagia, lethargy, de-
hydration, and severe body weight loss and was
DEQ-CFW 00001033

humanely sacrificed before the end of the study
(Butenhoff et al., 2002). The authors speculated,
based on a marked elevation of creatinine phospho-
kinase and marked decrease in serum cholesterol
observed in this monkey, that death may have been
due to lactic acidosis. Certain HMG CoA reductase
inhibitors (statins), especially when given in com-
bination with fibrates, have been shown to reduce
ubiquinone with resulting rhabdomyolysis and po-
tential lactic acidosis (De Pinieux et al., 1996; Flint
et al., 1997; Omar et al., 2001). The authors of this
study also speculated that the hepatic mitochondrial
proliferation observed in male monkeys treated with
PFOA may have been the result of downregulation
of HMG CoA reductase (Haughom and Spydevold,
1992) leading to a reduction in ubiquinone, de-
creased oxidative phosphorylation, and a concomi-
tant increase in mitochondrial synthesis (Fosslien,
2001).
Carcinogenicity
As discussed earlier, a 2-yr chronic rat bioas-
say with PFOA indicates a significantly increased
incidence of tumors (mainly adenomas) of the liver,
testis (Leydig cell), and pancreas (acinar cell)
(Biegel et al., 2001). This is consistent with the fact
that PFOA is a PP, since although most attention
has been focused on PP -induced liver tumors,
both Leydig-cell tumors (Cook et al., 1999) and
pancreatic acinar-cell tumors (Reddy and Rao,
1977; Svoboda and Azamoff, 1979) are often
observed following chronic exposure of rodents to
other PPs. Ohmura et al. (1997) showed that the per-
oxisome proliferator 4-chloro-6-(2,3-xylidino)-2-
pyrimidinylthio-(N-beta-hydroxyethyl)acetamide
(DR931), a potent hepatocarcinogen in rats, was
capable of inducing DNA synthesis in pancreatic
acinar cells, yet had no effect on ductal or islet
cells. Induction of liver, Leydig-cell, and pancreatic
acinar-cell tumors is a common finding for PPs
(Cook et al., 1999). In chronic bioassays in rats,
Cook et al. (1999) reported that 7 out of 11 PPs
induced all 3 tumor types (Cook et al., 1999), and 10
of the 11 PPs produced liver and Leydig-cell tumors
(Cook et al., 1999). The presence and established
persistence of PFOA in human tissues combined
with the increased incidence of liver, Leydig-cell,
and pancreatic acinar-cell tumors in chronic studies
with PFOA in rats raises questions concerning the
potential relevance of tumors observed in chronic
studies in rats to potential human health risk.
The liver tumors observed in rats after chronic
dietary exposure to PFOA are believed to have re-
sulted from peroxisome proliferation (Biegel et al.,
2001). Evidence supporting peroxisomal prolifera-
tion as a mechanism for PFOA-induced liver tumors
in rats comes from the measurement of hepatocellu-
lar peroxisome proliferation at 3-mo intervals dur-
ing the study by Biegel et al. (2001). Increased liver
weights and hepatic 8-oxidation activity were ob-
served in the PFOA-treated rats at all time points;
however, PFOA did not significantly increase hep-
atic cell proliferation.
Several mechanisms have been proposed for the
induction of hepatocellular tumors in PP -treated ro-
dents (Ashby et al., 1994; IARC, 1995; Reddy and
Rao, 1995). Although not yet fully elucidated, these
mechanisms typically involve a combination of ox-
idative stress (reactive oxygen species from 1-12O2)
resulting from enhanced peroxisomal ,B-oxidation
of fatty acids and an increase in cell proliferation;
there is also evidence for the involvement of altered
cell differentiation and decreased apoptosis (Vanden
Heuvel, 1999). With respect to oxidative damage,
Handler et al. (1992) did not observe an increase
in H2O2 production in liver, yet Kawashima et al.
(1994) and Permadi et al. (1992) observed an in-
crease in peroxide detoxifying enzymes after PFOA
treatment, and Panaretikis et al. (2001) observed an
increase in reactive oxygen species in HepG2 cells
exposed to PFOA. Takagi et al. (1991) observed an
increase in 8-hydroxydeoxyguanosine adducts to rat
liver DNA; however, it should be noted that this
study may have been influenced by a lack of suffi-
cient antioxidant to prevent artefactual creation of
adducts.
It is generally agreed that liver tumors in rats
produced by PPs are unlikely to be relevant to hu-
mans (Bentley et al., 1993; Ashby et al., 1994;
Cattley et al., 1998; Doull et al., 1999). A large
number of humans have been treated for relatively
long periods of time with hypolipidemic drugs that
are potent PPs in rodents. No significant changes in
the peroxisome number or volume occur in humans
taking substantial doses of these drugs for extended
periods of time (up to 3 yr) (Ashby et al., 1994).
Therefore, rodents appear to be poor models for hu-
man risk assessment with respect to liver effects ob-
served with PPs. The reason for the nonresponsive-
ness of humans to PPs is not yet fully understood,
although research shows differences in amount and
expression of PPARa between humans and rodents
(Cattley et al., 1998; Palmer et al., 1998).
Experimental evidence for the mechanism of
PFOA-induced Leydig-cell tumor formation, while
373
DEQ-CFW 00001034

not conclusive, tends to support the hypothesis that
a sustained increase in estradiol within the testes
may be responsible for the increased incidence of
Leydig-cell tumors in male Sprague-Dawley rats
(Cook et at., 1992; Biegel et at., 1995; Liu et al.,
1996a, 1996b). It was initially thought that PP -
induced Leydig-cell tumors arose by a mechanism
similar to that suggested for liver (Cook et al.,1992).
In the chronic PFOA study, however, there was no
evidence that peroxisomes were induced in Leydig
cells and there was no increase in P-oxidation activ-
ity above the normal baseline level (about 20 times
less than that in liver) (Biegel et al., 2001). Attention
has subsequently focused on the role in Leydig-cell
hyperplasia and neoplasia of a sustained increase
in the level of serum estradiol observed in PFOA-
treated rats (Liu et al., 1996a, 1996b; Biegel et al.,
1995, 2001). It has been proposed that PPs, includ-
ing PFOA, increase serum estradiol levels as well
as levels in the testis (interstitial fluid) via induction
of the enzyme aromatase, a CYP450-mediated en-
zyme in the liver (Liu et al., 1996b; Biegel et al.,
1995, 2001; Upham et al., 1998). It is then pro-
posed that the increase in testicular estradiol will
modulate growth factors, specifically transforming
growth factor (TGF)a, to stimulate cell proliferation
in the Leydig cell as it is known to do in other (e.g.,
mammary) tissues (Liu et al., 1987). The suggested
role of elevated estradiol in Leydig-cell neoplasia
is still uncertain because estrogenic compounds do
not induce Leydig-cell tumors in rats. It is possi-
ble, however, that the failure of estrogenic com-
pounds to cause Leydig-cell tumors results from a
depression of LH, which has been demonstrated to
be the chief "driver" of such tumors (Biegel et al.,
2001).
A second mechanism/pathway that has been
proposed to be also involved in the formation of
PFOA-induced Leydig-cell tumors is the inhibition
of testosterone biosynthesis, which disrupts the HPT
axis. In ex vivo studies with PFOA, exposure to
PFOA resulted in an inhibition of enzymes criti-
cal to the testosterone biosynthetic pathway, which,
if occurring in vivo, would subsequently lead to a
decrease in circulating testosterone (Biegel et al.,
1995). In studies in vitro, 13 PPARa agonists were
demonstrated to inhibit testosterone biosynthesis
(Liu et al., 1996a). The decrease in testosterone
levels results in compensatory increases in LH, in-
creased binding of LH to the LH receptor on Leydig
cells, and thus increases in Leydig-cell proliferation
(Clegg et al., 1997). In a 2-yr mechanistic bioassay
with PFOA and another PPARa agonist, LH levels
were not consistently increased, yet estradiol levels
374
were increased at several time points (Biegel et al.,
2001). This is consistent with PFOA inducing cir-
culating estradiol levels, which would attenuate the
elevation of LH caused by the inhibition of testos-
terone biosynthesis.
The extent to which the occurence of Leydig-
cell tumors in rats may be linked to PPARa act-
ivation is not clear. Other PPs (DEHP and clofibrate)
have been shown to increase serum estradiol con-
centrations in male rats (Eagon et al., 1994, 1996),
and several PPs (e.g., clofibrate, DEHP, gemfibrozil,
dibutyl phthalate, and Wyeth 14,643) have been
shown to reduce estradiol metabolism, resulting in
an increase in circulating levels of estradiol (Corton
et al., 1997; Eagon et al., 1994, 1996; Fan et al.,
1998). This pattern of hormonal alteration has also
been observed in vitro, where 10 of 11 peroxisome
proliferators evaluated increased estradiol levels,
and 11 of these PPs decreased testosterone levels
(Liu et al., 1996a, 1996b). While most PPs may in-
crease estradiol levels in rats, the direct association
of elevated estradiol with the production of Leydig-
cell tumors remains to be demonstrated. There are
seven proposed mechanisms for Leydig-cell tumori-
genesis in rodents, all of which disrupt the hormonal
milieu within the testes (Clegg et al., 1997; Cook
et al., 1999). The attribution of sustained estradiol
increase as part of the response to PPARa activa-
tion and as the operative mechanism for PFOA-
induced Leydig-cell tumors as well as the relevance
of these tumors to humans will require additional
research.
There is, however, evidence indicating a sub-
stantial difference in the susceptibility of rats and
humans to Leydig-cell tumorigenesis. There are nu-
merous pharmaceuticals and chemicals that have
been documented to produce Leydig-cell tumors in
rats and other laboratory animals, but not in hu-
mans. These include androgen -receptor antagonists,
dopamine agonists, estrogen agonists/antagonists,
other PPAR agonists (clofibrate, gemfibrozil), sug-
ars (lactose, lactitol), and nicotine (Clegg et al.,
1997). Studies have also indicated that there does
not appear to be a difference in the morphology of
Leydig-cell tumors, whether spontaneous or chem-
ically induced.
Since PFOA and other PPs do not increase
peroxisome levels in the testis, current ideas re-
garding the mechanism of PP -induced Leydig-cell
hyperplasia and neoplasia suggest the involvement
of hormone -mediated (estradiol) induction of tes-
ticular growth factors (Biegel et al., 2001). The
increased levels of estradiol are thought to result
from the induction of a hepatic CYP450-mediated
DEQ-CFW 00001035

aromatase. If this is not part of a pleiotropic
PPAR-mediated response, Leydig-cell tumorigenic-
ity could involve a distinct hormone -mediated
mechanism that might be of some relevance to hu-
mans. There is, however, evidence indicating a sub-
stantial difference in the susceptibility of rats and
humans to Leydig-cell tumorigenesis. Thus, while
the spontaneous incidence of Leydig-cell adeno-
mas in ageing Crl:CD BR rats ranges from approx-
imately 0 to 12% and can approach 100% in F344
rats, the rate in humans is reported to be only about
0.4 per million (0.00004%) (Schottenfeld, 1996).
It seems highly unlikely that PFOA exposures rep-
resent any significant human risk with respect to
Leydig-cell cancer. There was no substantial evi-
dence of any increase in serum estradiol or testic-
ular cancer in 3M plant workers exposed to PFOA
(Olsen et al., 1998), and in the 6-mo PFOA study
with cynomolgus monkeys, estradiol levels were
not increased (Butenhoff et al., 2002). Furthermore,
there was no increased incidence of testicular or
other cancers in humans ingesting high daily doses
of fibrates and other rodent PPs for hyperlipidemia
(Ashby et al., 1994).
The mechanism by which PFOA and some
other PPs induce pancreatic acinar-cell tumors is
not well understood. It has been hypothesized that it
might involve release of CCK in the gut with subse-
quent stimulation of the acinar cells in the pancreas
to secrete pancreatic enzymes into the gut (Biegel
et al., 2001; Herrington and Adrian, 1995). How-
ever, it must be concluded that, at the present time,
this is a speculative mechanism that is not supported
by experimental evidence for PFOA (Biegel et al.,
2001; Butenhoff et al., 2002). Even if this is found
to be the mechanism operating in rats, its applica-
bility to humans is highly uncertain (Gavin et al.,
1996, 1997; Cattley et al., 1998; Pandol, 1998). To
assess the hypothesis in production workers at 3M,
plasma CCK-33 measurements were made by ra-
dioimmunoassay during the 1997 medical surveil-
lance of 75 male production workers (Olsen et al.,
2000). The results showed a weak negative associa-
tion between serum PFOA and plasma CCK levels
and serum liver enzyme tests showed no indication
of cholestasis. There was no statistically significant
increased mortality from pancreatic cancer in these
workers (Alexander et al., 2002).
Since pancreatic acinar-cell adenomas are rare
in humans (Anderson et al., 1996), and since the rel-
evance of the rat PFOA data to humans is highly un-
certain, there is no reason at this time to believe that
PFOA represents a human health risk for pancreatic
cancer. This specific issue has been addressed by a
workshop at the International Life Sciences Institute
(Klaunig et al., 2003), and the results of the work-
shop are in press. When considering the relevance
of PFOA-induced pancreatic acinar-cell tumors in
rats to human health risk, the nongenotoxic mech-
anism (with a likely threshold) and the relatively
low exposure in human general populations should
be taken into account. (For a discussion of general
population risk, see Butenhoff et al., 2004a.)
CONCLUSIONS
Perfluorooctanoic acid is a fully fluorinated car-
boxylic acid that is used primarily as the ammonium
salt to aid in the emulsion polymerization of fluo-
ropolymers. Perfluorooctanoic acid and its salts are
soluble in water and readily dissociate to the car-
boxylate anion, perfluorooctanoate (PFOA). As a
result of the presence and biopersistence of PFOA
in the blood of humans, the potential health effects
of PFOA have been extensively examined. PFOA is
efficiently absorbed following oral, dermal, or in-
halation exposure. It is not metabolized and is elim-
inated intact. Excretion in urine appears to be the
primary route of elimination, and there is evidence
of biliary excretion and enterohepatic recirculation.
The rate of clearance of PFOA differs markedly be-
tween species and between sexes of a single species.
It is distributed mainly in the liver, serum and kid-
neys. PFOA is a mild skin irritant and moderate
eye irritant and is not a dermal sensitizing agent.
PFOA exhibits moderate acute oral and inhalation
toxicity and slight acute dermal toxicity. Signs and
symptoms of toxicity include body weight loss, liver
weight increase, and liver effects as demonstrated by
increased serum transaminase activity and diffuse
hepatocellular hypertrophy accompanied, at higher
doses, by acidophilic degeneration and/or necrosis
of the liver. Therefore, the liver appears to be the pri-
mary target organ. PFOA exhibited no teratogenic or
fetotoxic effects at doses below those causing mater-
nal toxicity and there is no evidence of any adverse
effects on reproductive success in a two -generation
reproduction study. Two-year cancer bioassays in
rats resulted in increased incidence of benign tu-
mors (adenomas) of the liver, pancreas (acinar cell)
and testes (Leydig cell) at 300 ppm in the diet, but
not at 30 ppm. A battery of tests for genotoxicity
demonstrates that PFOA does not cause either point
mutations or chromosomal aberrations.
375
DEQ-CFW 00001036

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