REVISED
DRAFT
HAZARD
ASSESSMENT
OF
PERFLUOROOCTANOIC
ACID
AND
ITS
SALTS
U.
S.
Environmental
Protection
Agency
Office
of
Pollution
Prevention
and
Toxics
Risk
Assessment
Division
November
4,
2002
PREFACE
This
is
a
preliminary
assessment
of
the
potential
hazards
to
human
health
and
the
environment
associated
with
exposure
to
perfluorooctanoic
acid
(
PFOA)
and
its
salts.
The
majority
of
the
toxicology
information
is
for
ammonium
perfluorooctanoic
acid
(
APFO).
This
assessment
includes
a
review
of
the
studies
that
were
available
as
of
October,
2002.
Table
of
Contents
Executive
Summary
1
1.0
Chemical
Identity
7
1.1
Physicochemical
Properties
7
2.0
Production
of
PFOA
and
its
Salts
9
2.1
Uses
of
PFOA
and
its
Salts
11
2.2
Environmental
Fate
12
2.2.1
Photolysis
12
2.2.2
Volatility
12
2.2.3
Biodegradation
13
2.2.4
Hydrolysis
13
2.2.5
Bioaccumulation
14
2.2.6
Soil
Adsorption
15
2.3
Environmental
Exposure
15
2.3.1
Discharge
to
Air
15
2.3.2
Discharge
to
Water
15
2.3.3
Discharge
to
Land
16
2.3.4
Environmental
Monitoring
16
2.4
Human
Biomonitoring
18
3.0
Human
Health
Hazards
23
3.1.
Metabolism
and
Pharmacokinetics
23
3.1.1
Half­
life
in
Humans
23
3.1.2
Absorption
Studies
in
Animals
24
3.1.3
Distribution
Studies
in
Animals
25
3.1.4
Metabolism
Studies
in
Animals
28
3.1.5
Elimination
Studies
in
Animals
28
3.2
Epidemiology
Studies
32
3.2.1
Medical
Surveillance
Studies
from
Antwerp
and
Decatur
Plants
32
3.2.2
Medical
Surveillance
Studies
from
Cottage
Grove
Plant
34
3.2.3
Mortality
Studies
36
3.2.4
Hormone
Study
39
3.2.5
Study
on
Episodes
of
Care
(
Morbidity)
40
3.3
Acute
Toxicity
Studies
in
Animals
42
3.3.1
Oral
Studies
42
3.3.2
Inhalation
Studies
43
3.3.3
Dermal
Studies
43
3.3.4
Eye
Irritation
Studies
43
3.3.5
Skin
Irritation
Studies
44
3.4
Mutagenicity
Studies
44
3.5
Subchronic
Toxicity
Studies
in
Animals
44
3.6
Developmental
Toxicity
Studies
in
Animals
60
3.7
Reproductive
Toxicity
Studies
in
Animals
64
3.8
Carcinogenicity
Studies
in
Animals
74
3.8.1
Cancer
Bioassays
74
3.8.2
Mode
of
Action
Studies
75
3.8.2.1
Liver
Tumors
76
3.8.2.2
Leydig
Cell
Tumors
76
3.8.2.3
Mammary
Gland
Tumors
77
3.8.2.4
Pancreatic
Tumors
77
3.9
Immunotoxicology
Studies
in
Animals
78
4.0
Hazards
to
the
Environment
80
4.1
Introduction
80
4.2
Acute
Toxicity
to
Freshwater
Species
81
5.0
References
87
ANNEX
I
 
Robust
Summaries
104
1
EXECUTIVE
SUMMARY
Introduction
Perfluorooctanoic
acid
(
PFOA)
and
its
salts
are
fully
fluorinated
organic
compounds
that
can
be
produced
synthetically
or
through
the
degradation
or
metabolism
of
other
fluorochemical
products.
PFOA
is
primarily
used
as
a
reactive
intermediate,
while
its
salts
are
used
as
processing
aids
in
the
production
of
fluoropolymers
and
fluoroelastomers
and
in
other
surfactant
uses.
In
recent
years,
less
than
600
metric
tons
per
year
of
PFOA
and
its
salts
have
been
manufactured
in
the
United
States
or
imported.
Most
of
the
toxicology
studies
have
been
conducted
with
the
ammonium
salt
of
perfluorooctanoic
acid,
which
is
referred
to
as
APFO
in
this
report.

Environmental
Fate
and
Effects
PFOA
is
persistent
in
the
environment.
It
does
not
hydrolyze,
photolyze
or
biodegrade
under
environmental
conditions.

Groundwater
samples
taken
near
fire­
training
areas
that
used
fire­
fighting
foams
containing
perfluorinated
surfactants
had
elevated
PFOA
concentrations
many
years
after
the
foam
use.
This
demonstrates
the
following:
(
1)
PFOA
either
existed
in­­
or
was
formed
via
degradation
of­­
the
surfactants,
(
2)
PFOA
or
its
precursors
migrate
through
the
soil,
and
(
3)
PFOA
persists
in
groundwater.

Several
wildlife
species
have
been
sampled
around
the
world
to
determine
levels
of
PFOA.
PFOA
has
rarely
been
found
in
fish
sampled
from
the
U.
S.,
certain
European
countries,
the
North
Pacific
Ocean
and
Antarctic
locations,
or
in
fish­
eating
bird
samples
collected
from
the
U.
S.,
including
Midway
atoll,
the
Baltic
and
Mediterranean
Seas,
and
Japanese
and
Korean
coasts.
PFOA
was
found
in
a
few
mink
livers
from
Massachusetts,
but
not
found
in
mink
from
Louisiana,
South
Carolina
and
Illinois.
PFOA
concentrations
in
river
otter
livers
from
Washington
and
Oregon
States
were
less
than
the
quantification
limit
of
36
ng/
g,
wet
wt.
PFOA
was
not
detected
at
quantifiable
concentrations
in
oysters
collected
in
the
Chesapeake
Bay
and
Gulf
of
Mexico
of
the
U.
S.
coast.

The
concentrations
of
PFOA
in
surface
water,
sediments,
clams,
and
fish
collected
from
locations
upstream
and
downstream
of
the
3M
manufacturing
facility
at
Decatur
AL
have
been
determined.
Of
the
three
downstream
water
and
sediment
sampling
locations,
the
two
closest
to
the
facility
had
PFOA
surface
water
concentrations
significantly
greater
than
the
two
upstream
sites;
the
three
downstream
locations
also
had
sediment
concentrations
significantly
greater
than
the
upstream
sites.
The
small
sample
size
prevented
determination
of
significance
for
fish
whole
body
PFOA
concentrations.
The
average
PFOA
concentration
in
clams
was
not
significantly
different
between
the
upstream
and
downstream
locations.

Based
on
available
laboratory
data,
APFO
does
not
appear
to
bioaccumulate
in
fish.
In
a
study
of
fathead
minnows,
the
calculated
BCF
for
APFO
was
1.8.
In
a
study
of
carp,
the
BCF
ranged
2
from
3.1
to
9.1.

Several
species
were
tested
to
assess
the
acute
toxicity
of
APFO;
these
included
the
fathead
minnow
(
Pimephales
promelas),
bluegill
sunfish
(
Lepomis
machrochirus),
water
flea
(
Daphnia
magna),
and
a
green
algae
(
Selenastrum
capricornutum).
Comparisons
of
the
different
studies
are
problematic
for
several
reasons.
The
studies
were
conducted
with
different
test
substances.
Generally
the
ammonium
salt
or
the
tetrabutylammonium
salt
was
tested.
Purity
of
the
test
material
is
a
major
concern
and
was
not
sufficiently
characterized
in
these
tests.
In
some
tests
it
appeared
that
100%
test
chemical
was
used,
for
others
a
chemical
of
lesser
purity
(
approximately
27
to
85%)
was
used.
Water,
a
solvent
(
isopropanol)
or
a
combination
of
both
was
used
in
other
tests,
for
no
obvious
stated
reason.
Finally,
only
nominal
test
chemical
concentrations
were
reported;
the
actual
concentrations
were
not
reported.

Twelve
tests
were
conducted
with
fathead
minnows;
96­
h
LC50
values
(
based
on
mortality)
ranged
from
70
to
843
mg/
L.
It
is
unclear
why
this
range
is
so
wide.
Assuming
these
studies
are
valid,
and
due
to
the
limitations
discussed
above,
these
toxicity
values
indicate
low
toxicity.
The
two
acute
values
for
bluegill
sunfish
also
indicate
low
toxicity
(
96­
h
LC50s
of
>
420,
and
569
mg/
L).

Nine
acute
tests
were
conducted
with
daphnids
and
48­
h
EC50
values
(
based
on
immobilization)
ranged
from
39
to
>
1000
mg/
L.
The
lower
values
are
indicative
of
moderate
toxicity,
but
the
wide
range
makes
interpretation
difficult.

Seven
tests
were
conducted
with
green
algae;
96­
h
EC50
values
(
based
on
growth
rate,
cell
density,
cell
counts,
and
dry
weights)
ranged
from
1.2
to
>
666
mg/
L
(
the
EC50
cell
density
value
of
1,000
mg/
L
is
excluded
from
this
discussion).
The
lower
value
indicates
high
to
moderate
toxicity,
based
on
the
acute
criteria.
The
lower
value
would
also
be
indicative
of
moderate
toxicity,
based
on
the
chronic
moderate
criterion
(
0.1<
10
mg/
L).
A
14­
d
EC50
value
of
43
mg/
L,
based
on
cell
counts,
for
green
algae
was
also
calculated
in
one
study.
This
is
indicative
of
low
chronic
toxicity,
based
on
the
chronic
criterion
(
10
mg/
L).
Green
algae
appeared
to
be
the
most
sensitive
test
species
in
the
44%
APFO
test
sample,
daphnids
were
the
next
most
sensitive,
and
fathead
minnows
were
the
least
sensitive.

Human
Health
Effects
and
Biomonitoring
Little
information
is
available
concerning
the
pharmacokinetics
of
APFO
in
humans.
An
ongoing
5­
year,
half­
life
study
in
9
retired
workers
has
suggested
a
mean
serum
PFOA
half­
life
of
4.37
years
(
range,
1.50
 
13.49
years).
These
data
provide
evidence
of
the
potential
to
bioaccumulate
PFOA
in
humans.

Animal
studies
have
shown
that
APFO
is
well
absorbed
following
oral
and
inhalation
exposure,
and
to
a
lesser
extent
following
dermal
exposure.
In
the
past,
Chemolite
workers
have
been
exposed
to
large
dermal
doses
of
PFOA.
It
appears
that
dermal
exposure
may
have
played
a
significant
role
in
the
absorption
of
PFOA
in
these
workers.
Upon
recognition
that
PFOA
could
3
be
absorbed
dermally,
work
practices
were
changed
and
engineering
controls
were
adopted
that
reduced
dermal
exposures.

PFOA
distributes
primarily
to
the
liver,
plasma,
and
kidney,
and
to
a
lesser
extent,
other
tissues
of
the
body
including
the
testis
and
ovary.
It
does
not
partition
to
the
lipid
fraction
or
adipose
tissue.
PFOA
binds
to
macro­
molecules
in
the
tissues
listed
above.
PFOA
is
not
metabolized
and
there
is
evidence
of
enterohepatic
circulation
of
the
compound.
The
urine
is
the
major
route
of
excretion
of
PFOA
in
the
female
rat,
while
the
urine
and
feces
are
both
major
routes
of
excretion
in
male
rats.
There
are
major
gender
differences
in
the
elimination
of
PFOA
in
rats.
In
female
rats,
the
half­
life
is
24
h
in
the
serum
and
60
h
in
the
liver;
in
male
rats,
the
half­
life
is
105
h
in
the
serum
and
210
h
in
the
liver.
The
rapid
excretion
of
PFOA
by
female
rats
is
due
to
active
renal
tubular
secretion
(
organic
acid
transport
system);
this
renal
tubular
secretion
is
believed
to
be
hormonally
controlled,
since
castrated
male
rats
treated
with
estradiol
have
excretion
rates
of
PFOA
similar
to
those
of
female
rats.
Hormonal
changes
during
pregnancy
do
not
appear
to
change
the
rate
of
elimination
in
rats.
This
gender
difference
has
not
been
observed
in
primates
and
humans.

There
are
limited
data
on
PFOA
serum
levels
in
workers
and
the
general
population.
Occupational
data
from
plants
in
the
U.
S.
and
Belgium
that
manufacture
or
use
PFOA
indicate
that
the
most
recent
mean
serum
levels
in
workers
range
from
0.84
to
6.4
ppm.
The
highest
level
reported
in
a
worker
in
1997
was
81.3
ppm.
In
non­
occupational
populations,
serum
PFOA
levels
were
much
lower.
In
both
pooled
blood
bank
samples
and
in
individual
samples,
mean
serum
PFOA
levels
ranged
from
3
to
17
ppb.
The
highest
serum
PFOA
levels
were
reported
in
a
sample
of
children
from
different
geographic
regions
in
the
U.
S.
(
range,
1.9
 
56.1
ppb).

Epidemiological
studies
on
the
effects
of
PFOA
in
humans
have
been
conducted
on
workers.
Two
mortality
studies,
a
morbidity
study,
and
studies
examining
effects
on
the
liver,
pancreas,
endocrine
system,
and
lipid
metabolism,
have
been
conducted
to
date.
In
addition,
a
crosssectional
as
well
as
a
longitudinal
study
of
the
worker
surveillance
data
have
recently
been
submitted.
However,
these
latest
2
studies
focus
primarily
on
PFOS
rather
than
PFOA,
even
though
recent
PFOA
levels
are
similar
to
or
higher
than
PFOS
levels
in
workers
at
these
plants.
(
It
should
be
noted
that
PFOS
levels
in
the
sampled
general
population
are
much
higher
than
PFOA
levels).

A
retrospective
cohort
mortality
study
demonstrated
a
weak,
although
statistically
significant
association
between
prostate
cancer
mortality
and
employment
duration
in
the
chemical
facility
of
a
plant
that
manufactures
PFOA.
However,
in
a
recent
update
to
this
study
in
which
more
specific
exposure
measures
were
used,
a
significant
association
for
prostate
cancer
was
not
observed.
In
a
morbidity
study,
workers
with
the
highest
PFOA
exposures
for
the
longest
durations
sought
care
more
often
for
prostate
cancer
treatment
than
workers
with
lower
exposures.

Another
study
reported
an
increase
in
estradiol
levels
in
workers
with
the
highest
PFOA
serum
levels;
however,
none
of
the
other
hormone
levels
analyzed
indicated
any
adverse
effects.
Some
4
of
the
same
employees
who
participated
in
the
hormone
study
also
were
included
in
a
study
of
cholecystokinin
(
CCK)
levels
in
employees.
No
positive
association
was
noted
between
CCK
values
and
PFOA.
The
other
available
study
examined
cholesterol
and
other
serum
components
in
workers.
There
did
not
appear
to
be
any
significant
differences
among
workers
of
different
exposure
levels.
At
plants
where
the
serum
PFOA
levels
were
lower,
cross­
sectional
and
longitudinal
studies
found
positive
significant
associations
between
PFOA
and
cholesterol
and
triglyceride
levels.
In
addition,
a
positive,
significant
association
was
reported
between
PFOA
and
T3
hormone
and
a
negative
association
with
HDL
in
the
cross­
sectional
study.
There
are
many
limitations
to
these
studies,
and
therefore,
all
of
these
results
must
be
interpreted
carefully.

In
acute
toxicity
studies
in
animals,
the
oral
LD50
values
for
CD
rats
were
>
500
mg/
kg
for
males
and
250­
500
mg/
kg
for
females,
and
<
1000
mg/
kg
for
male
and
female
Wistar
rats.
There
was
no
mortality
following
inhalation
exposure
of
18.6
mg/
L
for
one
hour
in
rats.
The
dermal
LD50
in
rabbits
was
determined
to
be
greater
than
2000
mg/
kg.
APFO
is
a
primary
ocular
irritant
in
rabbits,
while
the
data
regarding
potential
skin
irritancy
are
conflicting.

APFO
is
not
mutagenic.
APFO
did
not
induce
mutation
in
either
S.
typhimurium
or
E.
coli
when
tested
either
with
or
without
mammalian
activation.
APFO
did
not
induce
gene
mutation
when
tested
with
or
without
metabolic
activation
in
the
K­
1
line
of
Chinese
hamster
ovary
(
CHO)
cells
in
culture.
APFO
did
not
induce
chromosomal
aberrations
in
human
lymphocytes
when
tested
with
and
without
metabolic
activation
up
to
cytotoxic
concentrations.
APFO
was
tested
twice
for
its
ability
to
induce
chromosomal
aberrations
in
CHO
cells.
In
the
first
assay,
APFO
induced
both
chromosomal
aberrations
and
polyploidy
in
both
the
presence
and
absence
of
metabolic
activation.
In
the
second
assay,
no
significant
increases
in
chromosomal
aberrations
were
observed
without
activation.
However,
when
tested
with
metabolic
activation,
APFO
induced
significant
increases
in
chromosomal
aberrations
and
in
polyploidy.
APFO
was
negative
in
a
cell
transformation
assay
in
C3H
10T
½
mouse
embryo
fibroblasts
and
in
the
mouse
micronucleus
assay.

Subchronic
studies
in
rats
and
mice
with
28
and
90­
days
of
exposure
have
demonstrated
that
the
liver
is
the
primary
target
organ.
In
rat
studies,
males
are
far
more
sensitive
than
females.
Dietary
exposure
to
APFO
for
90
days
resulted
in
significant
increases
in
liver
weight
and
hepatocellular
hypertrophy
in
female
rats
at
1000
ppm
(
76.5
mg/
kg/
day)
and
in
male
rats
at
doses
as
low
as
100
ppm
(
5
mg/
kg/
day).
Analyses
of
serum
and
liver
levels
of
PFOA
showed
a
marked
gender
difference
that
accounts
for
the
difference
in
sensitivity.
Chronic
dietary
exposure
of
rats
to
300
ppm
APFO
(
14.2
and
16.1
mg/
kg/
day
for
males
and
females,
respectively)
for
2
years
resulted
in
increased
liver
and
kidney
weights,
hematological
effects
and
liver
lesions
in
males
and
females.
In
addition,
testicular
masses
were
observed
in
males
at
300
ppm
and
ovarian
tubular
hyperplasia
was
observed
in
females
after
exposure
to
30
ppm
(
1.6
mg/
kg/
day),
the
lowest
dose
tested.

In
a
90­
day
study
with
rhesus
monkeys,
exposure
to
doses
of
30
mg/
kg/
day
or
higher
resulted
in
death,
lipid
depletion
in
the
adrenals,
hypocellularity
of
the
bone
marrow,
and
moderate
atrophy
of
the
lymphoid
follicles
in
the
spleen
and
lymph
nodes.
Unlike
rodent
studies,
analyses
of
the
5
serum
and
liver
levels
did
not
reveal
a
gender
difference
in
monkeys,
but
the
sample
size
was
very
small
(
N=
2).
In
a
6­
month
study
of
male
cynomolgus
monkeys,
dosing
of
animals
in
the
30
mg/
kg/
day
dose
group
was
stopped
from
days
11
 
21
because
of
toxicity.
When
dosing
was
resumed
on
day
22,
animals
received
20
mg/
kg/
day
and
this
group
was
designated
the
30/
20
mg/
kg/
day
group.
This
treatment
was
also
not
tolerated
and
treatment
was
stopped
for
3/
6
monkeys.
Mortality
was
observed
in
one
monkey
at
3
mg/
kg/
day
and
at
30/
20
mg/
kg/
day.
There
were
no
consistent
effects
on
hormone
levels.
Increased
absolute
and
relative
liver
weights
were
noted
at
3,
10
and
30/
20
mg/
kg/
day.
While
there
was
no
evidence
of
peroxisome
proliferation,
there
was
evidence
of
mitochondrial
proliferation
suggesting
a
different
mode
of
action
than
observed
in
rats.
The
serum
levels
were
highly
variable
and
should
be
treated
with
caution.
On
day
9
of
treatment,
the
serum
levels
were
126
±
36.1
µ
g/
mL
in
the
3
mg/
kg/
day
group
and
1597
±
2392
µ
g/
mL
in
the
30/
20
mg/
kg/
day
group,
and
during
weeks
26/
27,
the
serum
levels
were
52.5
±
9.14
µ
g/
mL
in
the
3
mg/
kg/
day
group
and
51.5
±
77.6
µ
g/
mL
in
the
30/
20
mg/
kg/
day
group.
The
LOAEL
for
this
study
was
3
mg/
kg/
day
and
a
NOAEL
was
not
established.

PFOA
is
immunotoxic
in
mice.
Feeding
C57Bl/
6
mice
a
diet
containing
0.02%
PFOA
resulted
in
adverse
effects
to
both
the
thymus
and
spleen.
In
addition,
this
feeding
regimen
resulted
in
suppression
of
the
specific
humoral
immune
response
to
horse
red
blood
cells,
and
suppression
of
splenic
lymphocyte
proliferation
in
response
to
LPS
and
ConA.
The
suppressed
mice
recovered
their
ability
to
generate
a
humoral
immune
response
when
they
were
fed
a
diet
devoid
of
PFOA.
Studies
using
transgenic
mice
showed
that
the
peroxisome
proliferator­
activated
receptor
alpha
was
involved
in
causing
the
adverse
effects
to
the
immune
system.

Prenatal
developmental
toxicity
studies
in
rats
resulted
in
death
and
reduced
body
weight
in
dams
exposed
to
oral
doses
of
100
mg/
kg/
day
or
by
inhalation
to
25
mg/
m3
APFO.
There
was
no
evidence
of
developmental
toxicity
after
oral
exposure
to
doses
as
high
as
150
mg/
kg/
day,
while
inhalation
exposure
to
25
mg/
m3
resulted
in
reduced
fetal
body
weights.
In
a
rabbit
oral
developmental
toxicity
study
there
was
a
significant
increase
in
skeletal
variations
after
exposure
to
50
mg/
kg/
day
APFO.
There
was
no
evidence
of
maternal
toxicity
at
50
mg/
kg/
day,
the
highest
dose
tested.

In
a
two­
generation
reproductive
toxicity
study
in
rats
exposed
to
0,
1,
3,
10,
and
30
mg/
kg/
day
APFO,
significant
increases
in
absolute
and
relative
liver
and
kidney
weights
were
observed
in
F0
males
at
1
mg/
kg/
day,
while
significant
reductions
in
absolute
and
relative
kidney
weights
were
observed
in
F0
females
at
30
mg/
kg/
day.
Reproductive
indices
were
not
affected
in
the
F0
animals.
Serum
levels
of
the
10
and
30
mg/
kg/
day
groups
were
measured
for
F0
males
after
mating
and
F0
females
at
weaning
of
the
F1
pups.
In
F0
males,
the
serum
levels
were
51.1+
9.30
and
45.3+
12.6
ug/
l,
respectively
for
the
10
and
30
mg/
kg/
day
groups,
and
in
F0
females,
the
serum
levels
were
0.37+
0.0805
and
1.02+
0.425
ug/
l,
respectively
for
the
10
and
30
mg/
kg/
day
groups.
In
F1
females,
there
was
a
significant
increase
in
post
weaning
mortality,
a
significant
decrease
in
mean
body
weight,
and
a
significant
delay
in
sexual
maturation
at
30
mg/
kg/
day.
In
F1
males,
significant
decreases
in
body
weights
and
body
weight
gains,
and
significant
changes
in
absolute
liver
and
spleen
weights
and
in
the
ratios
of
liver,
kidney,
and
spleen
weights­
to­
6
brain
weights
were
observed
in
all
treated
groups.
The
increase
in
post
weaning
mortality
and
the
delay
in
sexual
maturation
was
also
noted
in
F1
males
at
30
mg/
kg/
day.
Reproductive
indices
were
not
affected
in
the
F1
animals.
The
LOAEL
for
the
F1
females
was
30
mg/
kg/
day,
and
the
NOAEL
was
10
mg/
kg/
day;
the
LOAEL
for
F1
males
was
1
mg/
kg/
day
and
a
NOAEL
was
not
determined.
The
difference
in
sensitivity
is
presumed
to
be
related
to
the
gender
difference
in
elimination
of
APFO.
No
treatment­
related
effects
were
observed
in
the
F2
generation.
However,
it
should
be
noted
that
the
F2
pups
were
sacrificed
at
weaning,
and
thus
it
was
not
possible
to
ascertain
the
potential
post­
weaning
effects
that
were
noted
in
the
F1
generation.

Carcinogenicity
studies
in
Sprague­
Dawley
(
CD)
rats
show
that
APFO
is
weakly
carcinogenic,
inducing
Leydig
cell
adenomas
in
the
male
rats
and
mammary
fibroadenomas
in
the
females
following
dietary
exposure
to
300
ppm
for
2
years
(
equivalent
to
14.2
mg/
kg/
day
in
males
and
16.1
mg/
kg/
day
in
females).
APFO
has
also
been
reported
to
be
carcinogenic
toward
the
liver
and
pancreas
of
male
CD
rats
at
300
ppm.

The
mechanism(
s)
of
APFO
tumorigenesis
is
not
clearly
understood.
Available
data
appear
to
indicate
that
the
induction
of
tumors
by
APFO
is
due
to
a
non­
genotoxic
mechanism,
involving
activation
of
receptors
and
perturbations
of
the
endocrine
system.
There
is
sufficient
evidence
to
suggest
that
APFO
is
a
PPAR ­
agonist
and
that
the
liver
carcinogenicity/
toxicity
of
APFO
is
mediated
by
binding
to
PPAR 
in
the
liver.
Recently,
IARC
(
1995)
concluded
that
the
liver
tumors
induced
in
rodents
by
PPAR ­
agonists
are
unlikely
to
be
operative
in
humans
based
on
our
current
understanding
of
the
animal
mode
of
action.
The
Agency
is
currently
examining
the
scientific
knowledge
associated
with
PPAR ­
agonist
 
induced
liver
tumors
in
rodents
and
the
relevance
to
humans.
Available
data
suggest
that
the
induction
of
Leydig
cell
tumors
(
LCT)
and
mammary
gland
neoplasms
by
APFO
may
be
due
to
hormonal
imbalance
resulting
from
activation
of
the
PPAR 
and
induction
of
the
cytochrome
P450
enzyme,
aromatase.
Preliminary
data
suggest
that
the
pancreatic
acinar
cell
tumors
are
related
to
an
increase
in
serum
level
of
the
growth
factor,
cholecystokinin.
7
1.0
Chemical
Identity
Chemical
Name:
Perfluorooctanoic
Acid
Molecular
formula:
C8
H
F15
O2
Structural
formula:
F­
CF2­
CF2­
CF2­
CF2­
CF2­
CF2­
CF2­
C(=
O)­
X,

The
free
acid
and
some
common
derivatives
have
the
following
CAS
numbers:
The
perfluorooctanoate
anion
does
not
have
a
specific
CAS
number.

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

Ammonium
Salt
(
X
=
OM+;
M
=
NH4)
[
3825­
26­
1]
Sodium
Salt
(
X
=
OM+;
M
=
Na)
[
335­
95­
5]
Potassium
Salt
(
X
=
OM+;
M
=
K)
[
2395­
00­
8]
Silver
Salt
(
X
=
OM+;
M
=
Ag)
[
335­
93­
3]

Acid
Fluoride
(
X
=
F)
[
335­
66­
0]

Methyl
Ester
(
X
=
CH3)
[
376­
27­
2]
Ethyl
Ester
(
X
=
CH2­
CH3)
[
3108­
24­
5]

Synonyms:
1­
Octanoic
acid,
2,2,3,3,4,4,5,5,6,6,7,7,8,8,8­
pentadecafluoro­
PFOA
1.1
Physicochemical
Properties
For
this
report,
perfluorooctanoic
acid
is
consistently
referred
to
as
PFOA.
Most
of
the
toxicology
studies
have
been
conducted
with
the
ammonium
salt
of
perfluorooctanoic
acid,
which
will
be
referred
to
as
APFO
in
this
report.
PFOA
is
a
completely
fluorinated
organic
acid.
The
typical
structure
has
a
linear
chain
of
eight
carbon
atoms.
The
physical
chemical
properties
noted
below
are
for
the
free
acid,
unless
otherwise
stated.
The
data
for
the
free
acid,
pentadecafluorooctanoic
acid
[
335­
67­
1],
is
the
most
complete.
The
reported
vapor
pressure
of
10
mm
Hg
appears
high
for
a
low
melting
solid
when
compared
to
other
low
melting
solids
(
chloroacetic
acid:
solid;
MP
=
61
to
63
°
C;
BP
=
189
°
C;
VP
=
0.1
kPa
(
0.75
mm
Hg)
@
20
°
C;
NIOSH),
but
is
consistent
with
other
perfluorinated
compounds
with
similar
boiling
points
(
perfluorobutanoic
acid
BP
=
120
°
C,
VP
10
mm
Hg
@
20
°
C;
Beilstein
IV
2
p.
811).
Another
explanation
may
be
that
the
10
mm
vapor
pressure
was
measured
at
an
elevated
temperature
(
but
the
temperature
inadvertently
omitted),
as
perfluorooctanoic
acid
is
typically
handled
as
a
liquid
at
65
°
C
(
3M
data
sheet
for
FC­
26).

The
free
acid
is
expected
to
completely
dissociate
in
water,
leaving
the
anionic
carboxylate
in
the
water
and
the
perfluoroalkyl
chain
on
the
surface.
In
aqueous
solutions,
individual
molecules
of
PFOA
anion
loosely
associate
on
the
water
surface
and
partition
between
the
air
/
water
8
interface.
Several
reports
note
that
PFOA
salts
self­
associate
at
the
surface,
but
with
agitation
they
disperse
and
micelles
form
at
higher
concentrations.
(
Simister,
1992;
Calfours,
1985;
Edwards,
1997).
Water
solubility
has
been
reported
for
PFOA,
but
it
is
unclear
whether
these
values
are
for
a
microdispersion
of
micelles,
rather
than
true
solubility.
Due
to
these
same
surface­
active
properties
of
PFOA,
and
the
test
protocol
for
the
OECD
shake
flask
method,
PFOA
is
anticipated
to
form
multiple
layers
in
octanol/
water,
much
like
those
observed
for
PFOS.
Therefore,
an
n­
octanol/
water
partition
coefficient
cannot
be
determined.

The
available
physicochemical
properties
for
the
PFOA
free
acid
are:

MW:
414
(
Beilstein,
1975)
MP:
45
 
5
°
C
(
Beilstein,
1975)
BP:
189
 
192
°
C
/
736
mm
Hg
(
Beilstein,
1975)
VP:
10
mm
Hg
@
25
°
C
(
approx.)
(
Exfluor
MSDS)
Sol.
­
Water:
3.4
g/
L
(
telomeric
[
MP
=
34
°
C
ref.
0.01
­
0.02
mol/
L
~
4
­
8
g/
L)
(
MSDS
from
Merck,
Fischer,
and
Chinameilan
Internet
sites)
pKa:
2.5
(
USEPA
AR226­
0473)
pH
(
1g/
L):
2.6
(
MSDS
Merck)

The
PFOA
derivative
of
greatest
concern
and
most
wide
spread
use
is
the
ammonium
salt
(
APFO),
[
3825­
26­
1]).
This
substance
was
the
testing
substance
for
the
toxicity
testing
performed
and
has
the
P­
Chem
data
provided
in
the
Table
below.

The
water
solubility
of
APFO
has
been
inconsistently
reported.
One
3M
study
reported
the
water
solubility
of
APFO
to
be
>
10%.
It
was
noted
in
an
earlier
study
that
at
concentrations
of
20
g/
L,
the
solution
"
gelled"
(
3M,
1979).
These
numbers
seem
surprising
low
for
a
salt
,
in
light
of
Apollo
Scientific
selling
a
31%
aq.
solution
of
APFO.
One
author
reported
the
APFO
partition
coefficient
log
Pow
=
5.
Another
author
reported
an
estimated
APFO
log
Pow
=
­
0.9.
This
value
might
not
be
accurate
due
to
the
estimation
method
used
(
Hansch
and
Leo
1979).
Again,
the
anticipated
formation
of
an
emulsified
layer
between
the
octanol
and
water
surface
interface
would
make
determination
of
log
Kow
impossible.

Determination
of
the
vapor
pressure
of
APFO
is
complicated.
APFO
had
recently
reported
a
vapor
pressure
of
7
x
10­
5
mm
Hg
at
20E
C,
which
seems
too
low
for
a
material
that
sublimates
as
the
ammonium
salt.
(
3M
Environmental
Laboratory,
1993).
The
ammonium
salt
begins
to
sublimate
at
130
°
C
(
USEPA
AR­
226
473).
As
the
temperature
increases
from
when
APFO
begins
to
sublimate,
20%
of
the
sample
weight
is
lost
by
169
°
C.
Other
salts
(
Cs,
K,
Ag,
Pb,
Li)
do
not
demonstrate
similar
weight
loss
until
237
°
C
or
higher.
(
Lines,
1984).
Decomposition
of
different
salts
produces
perfluoroheptene
(
loss
of
metal
fluoride
and
carbon
dioxide).
This
occurs
at
320
°
C
for
the
sodium
salt
and
at
250­
290
°
C
for
the
silver
salt
(
Beilstein
1975).

The
physicochemical
properties
of
PFOA
and
its
common
derivatives
are
summarized
in
Table
1.
9
Table
1.
Reported
Physicochemical
Properties
Compound
CAS
REG
#
MP
BP
VP
Sol.­
H2O
Log
P*
Rf­
C(=
O)
F
335­
64­
8
131
°
C
Rf­
CO2H
335­
67­
1
55
°
C
189
°
C
10
mm
Hg
3.4
g/
L
Rf­
CO2­
NH4+
3825­
26­
1
130
°
C
(
sub)
sublimes
1
x
10E­
5
mm
Hg
20
g/
L
gels
Rf­
C(=
O)
OMe
376­
27­
2
159
°
C
pH
(
1
g
free
acid
/
L
Water)
=
1.5
 
2.5
Free
acid
pKa
is
approximately
0.6
Sodium
or
Silver
salts
of
PFOA
decompose
above
250
°
C
to
generate
perfluoroolefins.
 
Surfactants
traditionally
emulsify
octanol
and
water
2.0
Production
of
PFOA
and
its
Salts
PFOA
has
been
commercially
manufactured
by
two
major
alternative
processes:
1)
the
Simons
Electro­
Chemical
Fluorination
(
ECF)
process
or
2)
the
telomerization
process.
The
3M
Company
was
reported
to
be
largest
manufacturer
and
importer
of
PFOA
and
its
salts
in
the
United
States
in
1999.
They
predominantly
used
the
ECF
process
to
produce
a
PFOA
precursor
which
is
ultimately
converted
to
PFOA.
Data
for
the
FC­
26
(
97%
C8)
suggests
this
purified
material
was
about
80%
linear.

In
the
ECF
process
to
make
PFOA,
an
electric
current
is
passed
through
a
solution
of
anhydrous
hydrogen
fluoride
and
an
organic
feedstock,
typically
an
octanoic
acid
derivative.
The
ECF
process
replaces
the
carbon­
hydrogen
bonds
on
molecules
of
the
organic
feedstock
with
carbonfluorine
bonds,
in
an
identical
manner
used
to
make
PFOS.
Perfluorination
occurs
when
all
the
carbon­
hydrogen
bonds
are
replaced
with
carbon­
fluorine
bonds.
The
ECF
process
yields
between
30­
45
percent
straight
chain
(
normal)
perfluorooctanoyl
fluoride
(
PFOF),
along
with
a
variable
mixture
of
byproducts
and
impurities.
The
product
from
the
ECF
process
is
not
a
pure
chemical,
but
instead
a
mixture
of
isomers
and
homologs
including
higher
and
lower
straightchain
homologs;
branched­
chain
perfluoroalkyl
fluorides
of
various
chain
lengths;
straight­
chain,
branched,
and
cyclic
perfluroalkanes
and
ethers;
and
other
byproducts
(
3M
Company,
2000a).
After
separation
and
recovery
of
desired
material
from
the
byproducts
and
impurities
of
the
crude
reaction
mixture,
the
initially
formed
acid
fluoride
is
base
hydrolyzed
in
batch
reactors
and
acidified
to
ultimately
yield
PFOA.
The
PFOA
salts
are
synthesized
by
base
neutralization
of
the
acid
to
the
salt
in
a
separate
reactor
(
3M
Company,
2000b).

3M
has
characterized
its
manufacture
of
PFOA
and
its
derivatives
in
1997
at
less
than
500,000
kg
per
year
in
the
US,
and
its
importation
at
less
than
100,000
kg
(
3M
Company,
2000a).
(
These
figures
may
overstate
the
total
production
volume
of
PFOA
since
the
vast
majority
of
PFOA
is
consumed
in
the
manufacture
of
the
ammonium
or
other
metallic
salts.)
Industry
sources
have
characterized
3M
as
the
dominant
global
producer
of
PFOA­
related
chemicals,
manufacturing
approximately
85
percent
or
more
of
total
worldwide
volumes
of
the
ammonium
salt
of
PFOA
10
(
FMG,
2001).
USEPA
has
not
located
information
that
would
contradict
this
claim.
More
precise
production
volumes
of
PFOA
and
the
ammonium
and
sodium
salts
have
been
reported
to
USEPA
by
3M,
but
have
been
claimed
as
TSCA
confidential
business
information,
preventing
disclosure
in
this
report.
Since
1985,
USEPA
has
received
a
total
of
approximately
25
notifications
for
PFOA­
related
chemicals
that
were
not
previously
on
the
TSCA
Chemical
Inventory.
Most
of
these
notifications
were
from
companies
other
than
3M.
In
most
cases,
the
notifications
qualified
for
the
Low
Volume
Exemption
for
new
chemicals
with
a
production
volume
less
than
10
metric
tons
per
year.

Current
production
volume
information
for
manufacturers
other
than
3M
has
not
been
provided
by
industry,
nor
is
it
available
in
USEPA's
Chemical
Update
System
(
which
contains
information
on
non­
polymeric
organic
chemicals
manufactured
in
the
United
States
or
imported
in
volumes
above
4,525
kg).
Furthermore,
there
is
no
information
on
the
total
cumulative
production
volumes
of
PFOA
since
initial
commercialization.
In
terms
of
on­
going
production,
in
comments
to
the
draft
hazard
assessment
of
PFOA
3M
has
stated
that
their
May
16,
2000
announcement
of
the
phase­
out
of
the
production
of
perfluorooctanyl
chemistry
and
related
products
includes
PFOA
and
its
salts.
This
commits
3M
to
a
complete
phase­
out
of
PFOA
and
PFOA­
related
chemicals
identical
to
the
phase
out
of
PFOS
and
PFOS­
related
chemicals
by
the
end
of
2002.
In
addition,
the
Fluoropolymer
Manufacturers
Group
(
3M
largest
customer
for
PFOA
products,
specifically
the
ammonium
salt,
APFO)
has
informed
the
EPA
that
it
will
be
manufacturing
300,000
kg/
yr
to
meet
the
industry
demand
for
APFO
that
3M
has
not
met
(
FMG
2002
AR226­
1094).
The
FMG
manufacturing
process
is
based
on
telomerization.

The
telomerization
process,
according
to
3M's
Bultman
and
Pike,
"
is
the
reaction
of
a
telogen
(
such
as
pentafluoroethyl
iodide...)
with
a
polymerizable
ethylenic
compound
(
such
as
tetrafluoroethylene)
to
form
`
telomers'.
In
this
process
the
resultant
telomer
iodides
are
then
reacted
with
ethylene,
via
free
radical
addition.
This
forms
a
mixture
of
iodides
that
can
be
reacted
subsequently
to
form
a
variety
of
useful
materials....".
Dupont
uses
an
analogous
telomerization
process.
Either
process
yields
predominantly
pure
straight­
chain
acids
with
an
even
number
of
carbon
atoms.
Commercial
products
manufactured
through
the
telomerization
process
are
generally
mixtures
of
perfluorinated
compounds
with
ranges
of
even
carbon
numbers
(
Renner,
2001).
Distillation
can
be
used
to
obtain
pure
components
(
ECT,
1994).

Aside
from
the
United
States,
OECD
Member
countries
that
reportedly
have
production
capacity
include
France,
Germany,
Italy,
and
Japan.
There
may
also
be
some
production
in
non­
OECD
countries
such
as
China.
Following
are
companies
that
may
manufacture
PFOA
and
its
salts
(
3M
Company,
2000b;
Directory
of
World
Chemical
Producers,
1998;
Dynax,
2000;
Renner,
2001;
SEMI,
2001):

OECD
°
3M
Company
(
United
States)
°
DuPont
(
United
States)
°
Exfluor
Research
Corporation
(
United
States)
11
°
PCR
Inc.
(
United
States)
°
Ciba
Specialty
Chemicals
(
Germany)
°
Clariant
(
Germany)
°
Dyneon
(
Germany)
°
Hoechst
Aktiengesellschaft
(
Germany)
°
EniChem
Synthesis
S.
p.
A.
(
Italy)
°
Miteni
S.
p.
A.
(
Italy)
°
Asahi
Glass
(
Japan)
°
Daikin
(
Japan)
°
Dainippon
(
Japan)
°
Tohkem
Products
Corporation
(
Japan)

Non­
OECD
°
Chenguang
Research
Institute
of
the
Chemical
Industry
(
China)
°
Shanhai
3F
New
Materials
Co.,
Ltd.
(
China)

2.1
Uses
of
PFOA
and
its
Salts
PFOA
is
used
mainly
as
a
chemical
intermediate,
and
its
salts
are
used
in
emulsifier
and
surfactant
applications.

According
to
3M,
the
vast
majority
of
PFOA
is
consumed
to
make
the
ammonium
or
sodium
salts.
3M
also
uses
PFOA
as
a
reactive
intermediate
in
the
industrial
synthesis
of
a
fluoroacrylic
ester.
The
fluoroacrylic
ester
is
used
in
an
industrial
coating
application
(
3M
Company,
2000a).

The
salts
of
PFOA
have
additional
uses,
mostly
in
surfactant
and
emulsifier
applications.
These
include
the
following:

Processing
aid
in
the
industrial
synthesis
of
fluoropolymers
and
fluoroelastomers
such
as
polytetrafluoroethylene
and
polyvinylidene
fluoride
with
a
variety
of
industrial
and
consumer
uses
(
3M
Company,
2000a;
DuPont,
2000;
Daikin,
2001).

Post­
polymerization
processing
aids
in
the
stabilization
of
suspensions
of
fluoropolymers
and
fluoroelastomers
prior
to
further
industrial
processing
(
3M
Company,
2000a).

Processing
aid
for
factory­
applied
fluoropolymer
coatings
on
fabrics,
metal
surfaces,
and
fabricated
or
molded
parts
(
3M
Company,
2000a).

Extraction
agent
in
ion­
pair
reversed­
phased
liquid
chromatography
(
Petritis,
1999).

Based
on
the
physicochemical
properties
of
the
salts
of
PFOA,
they
may
also
have
other
related
surfactant
or
emulsifier
uses
as
a
photographic
chemical
or
in
the
manufacture
of
electronic
components
such
as
semiconductors.
These
same
properties
may
lead
industry
to
explore
PFOA
12
as
a
replacement
chemical
for
PFOS
in
other
applications
in
which
PFOA
is
not
currently
used.

2.2
Environmental
Fate
2.2.1
Photolysis
Direct
photolysis
of
APFO
was
examined
in
two
separate
studies
(
Todd,
1979;
Hatfield,
2001)
and
photodegradation
was
not
observed
in
either
study.
In
the
Todd
(
1979)
study,
a
solution
of
50
mg/
l
APFO
in
2.8
liters
of
distilled
water
was
exposed
to
simulated
sunlight
at
22
±
2
º
C.
Spectral
energy
was
characterized
from
290­
600
nm
with
a
max
output
at
~
360
nm.
Direct
photolysis
of
the
test
substance
was
not
detected.
However,
the
author
noted
that
sample
purity
was
not
properly
characterized
which
may
have
contributed
to
experimental
error.

In
the
Hatfield
(
2001)
study,
both
direct
and
indirect
photolysis
were
examined
utilizing
techniques
based
on
EPA
and
OECD
guidance
documents.
To
determine
the
potential
for
direct
photolysis,
APFO
was
dissolved
in
pH
7
buffered
water
and
exposed
to
simulated
sunlight
(
Scrano,
1999;
Nubbe,
1995).
For
indirect
photolysis,
APFO
was
dissolved
in
3
separate
matrices
and
exposed
to
simulated
sunlight
for
periods
of
time
from
69.5
to
164
hours.
These
exposures
tested
how
each
matrix
would
affect
the
photodegradation
of
APFO.
One
matrix
was
a
pH
7
buffered
aqueous
solution
containing
H2O2
as
a
well­
characterized
source
of
OH
radicals
(
Ogata,
1983;
Lunak,
1992).
This
tested
the
propensity
of
APFO
to
undergo
indirect
photolysis.
The
second
matrix
contained
Fe2O3
in
water
that
has
been
shown
to
generate
hydroxyl
radicals
via
a
Fenton­
type
reaction
in
the
presence
of
natural
and
artificial
sunlight
(
Kachanova,
1973;
Behar,
1966).
The
third
matrix
contained
a
standard
solution
of
humic
material.
Neither
direct
nor
indirect
photolysis
of
APFO
was
observed
based
on
loss
of
starting
material.
Predicted
degradation
products
were
not
detected
above
their
limits
of
quantitation.
There
was
no
conclusive
evidence
of
direct
or
indirect
photolysis
whose
rates
of
degradation
are
highly
dependent
on
the
experimental
conditions.
Using
the
iron
oxide
(
Fe2O3)
photoinitiator
matrix
model,
the
APFO
half­
life
was
estimated
to
be
greater
than
349
days.

2.2.2
Volatility
Impinger
studies
were
performed
to
examine
the
volatility
of
APFO
and
PFOS.
Solutions
of
APFO
or
PFOS
containing
ammonium
acetate
in
water/
1­
propanol
(
50:
50)
or
phase
transfer
agents,
e.
g.,
n­
alkyldimethylbenzylammonium
chloride
(
3M
Environmental
Laboratory,
1993)
were
blown
with
280
liters
of
air
at
a
flow
rate
of
1
L/
min.
(
3M
Environmental
Laboratory,
1993).
The
results
indicate
there
is
some
loss
of
APFO
and
PFOS,
but
most
of
the
solutions
retained
over
80%
or
more
of
the
fluorochemicals.
The
average
retention
was
92%
for
both
APFO
and
PFOS.
This
indicates
that
there
is
loss
from
the
solutions.
However,
some
of
the
solutions,
particularly
the
n­
alkyldimethylbenzylammonium
chloride
solution,
appear
to
retain
all
the
fluorochemicals.
These
results
were
reviewed
by
Dr.
Edwin
Tucker
of
the
Chemistry
Dept.
at
the
University
of
Oklahoma
(
3M
Environmental
Laboratory,
1993).
He
concluded
that
it
is
very
unlikely
that
these
fluorochemicals
were
removed
by
bubbling
air
through
water
due
to
13
their
very
low
vapor
pressures.
He
suggested
that
a
more
plausible
mechanism
for
loss
from
the
solution
phase
is
concentration
of
the
surfactants
in
foam
and
loss
from
the
bubbled
solutions
as
foam
or
micro­
droplets.

In
the
second
part
of
the
experiment,
air
was
passed
over
the
fluorochemicals
and
bubbled
through
a
train
of
impingers
containing
the
ammonium
acetate
solution.
It
was
expected
that
if
any
fluorochemicals
were
present
in
the
air
they
would
be
transferred
and
retained
by
the
ammonium
acetate
solution.
However,
no
fluorochemicals
were
present
in
either
the
first
or
second
impinger.
The
report
concludes
that
the
vapor
pressure
of
both
compounds
is
less
than
10
E­
07
mm
Hg.

According
to
these
experiments,
APFO
and
PFOS
(
potassium
salt)
have
very
low
volatility
and
vapor
pressure.
Quantitative
conclusions
regarding
rates
of
volatilization
from
water
or
Henry's
Law
constant
are
not
possible.
However,
APFO
and
PFOS
are
capable
of
transport
out
of
water.
Also,
the
loss
of
the
fluorochemicals
may
have
been
as
the
free
acids,
not
the
salt
forms.
APFO
sublimes
at
130
C
(
see
Physicochemical
Properties
Section
1.1).
There
is
no
information
on
the
validity
of
the
test
method
for
determining
volatility
of
the
test
substance.
The
study
also
lacks
characterization
of
the
purity
of
the
test
substance.

2.2.3
Biodegradation
Using
an
acclimated
sludge
inoculum,
the
biodegradation
of
APFO
was
investigated
using
a
shake
culture
study
modeled
after
the
Soap
and
Detergent
Association's
presumptive
test
for
degradation
(
Reiner,
1978).
Both
thin­
layer
and
liquid
chromatography
did
not
detect
the
presence
of
any
metabolic
products
over
the
course
of
2
1/
2
months
indicating
that
PFOA
does
not
readily
undergo
biodegradation.
In
a
related
study,
2.645
mg/
L
APFO
was
not
measurably
degraded
in
activated
sludge
inoculum
(
Pace
Analytical,
2001).
Test
flasks
were
prepared
using
a
mineral
salts
medium,
1
mL
methanol,
and
50
mL
settled
sludge.
Analysis
was
conducted
with
a
HPLC/
MSD
system.
Several
other
studies
conducted
between
1977­
1987
also
did
not
observe
APFO
biodegradation
using
what
probably
were
standard
COD
and
BOD
methods,
however,
the
methods
used
in
these
studies
were
either
insufficiently
described
(
i.
e.
no
description
of
experimental
protocols)
or
there
were
indications
of
a
high
degree
of
experimental
error.
The
results
were,
therefore,
deemed
unreliable
by
the
submitter
(
3M
Company,
1977;
3M
Company,
1980;
3M
Company,
1985b;
Pace
Analytical,
1997).

2.2.4
Hydrolysis
The
3M
Environmental
Laboratory
(
2001a)
performed
a
study
of
the
hydrolysis
of
PFOA.
The
study
procedures
were
based
on
EPA's
OPPTS
Guideline
Document
835.2110
(
EPA
1998);
although
the
procedures
do
not
fulfill
all
the
requirements
of
the
guideline,
they
were
more
than
adequate
for
these
studies.
Results
were
based
on
the
observed
concentrations
of
PFOA
in
buffered
aqueous
solutions
as
a
function
of
time.
The
chosen
analytical
technique
was
high
performance
liquid
chromatography
with
mass
spectrometry
detection
(
HPLC/
MS).
14
During
the
study,
samples
were
prepared
and
examined
at
six
different
pH
levels
from
1.5
to
11.0
over
a
period
of
109
days.
Experiments
were
performed
at
50
°
C
and
the
results
extrapolated
to
25
°
C.
Data
from
two
of
the
pH
levels
(
3.0
and
11)
failed
to
meet
the
data
quality
objective
and
were
rejected.
Also
rejected
were
the
data
obtained
for
pH
1.5
because
ion
pairing
led
to
artificially
low
concentrations
for
all
the
incubation
periods.
The
results
for
the
remaining
pH
levels
(
5.0,
7.0,
and
9.0)
indicated
no
clear
dependence
of
the
degradation
rate
of
PFOA
on
pH.
From
the
data
pooled
over
the
three
pH
levels,
it
was
estimated
that
the
hydrolytic
half­
life
of
PFOA
at
25
°
C
is
greater
than
92
years,
with
the
most
likely
value
of
235
years.
From
the
mean
value
and
precision
of
PFOA
concentrations,
it
was
estimated
the
hydrolytic
half­
life
of
PFOA
to
be
greater
than
97
years.

2.2.5
Bioaccumulation
Three
studies
have
been
conducted
to
determine
the
potential
for
bioaccumulation
of
APFO.
Howell
et
al.
(
1995)
exposed
Fathead
minnows
to
25
mg/
l
APFO
for
13
days.
After
13
days
exposure,
the
fish
were
then
removed
from
APFO
contaminated
water
and
analyzed
for
depuration
over
15
days.
After
192
and
312
hours
exposure
to
APFO
contaminated
water,
the
average
concentration
of
APFO
in
fish
tissue
was
44.7
and
46.7
µ
g/
g
wet
weight
(
ww),
respectively.
At
this
point,
APFO
appeared
to
reach
steady
state.
Twenty­
four
hours
after
being
transferred
to
clean
water,
the
concentration
of
APFO
decreased
to
19.9
µ
g/
g
ww
and
by
96
hours
post­
exposure,
the
concentration
had
decreased
to
approximately
8
µ
g/
g
ww
and
remained
relatively
constant
until
test
termination
at
360
hours.
The
calculated
BCF
for
APFO
was
1.8.
It
should
be
noted
that
questions
have
been
raised
about
this
study
regarding
the
analytical
techniques,
high­
test
chemical
concentration,
and
short
test
duration.

In
a
bioaccumulation
study
of
carp
(
Cyprinus
carpio),
28
carp
were
exposed
to
two
concentrations
of
PFOA,
5
and
50
:
g
/
L
respectively
for
28
days
in
a
flow
through
system
(
Kurume
Laboratory,
2001).
The
chemical
was
prepared
for
exposure
by
using
a
dispersant
made
up
of
hydrogenated
castor
oil
and
mixed
in
acetone.
The
purity
of
PFOA
was
100%.
Water
quality
was
monitored
daily
throughout
the
duration.
The
bioaccumulation
potential
of
PFOA
was
found
to
be
low.
The
carp
exposed
to
5
:
g/
L
resulted
in
a
BCF
of
3.1
while
the
carp
exposed
to
50
:
g
/
L
showed
a
BCF
of
<
5.1
 
9.1.

Vraspir
(
1979)
conducted
a
study
to
determine
if
bluegill
sunfish
bioaccumulate
fluorochemicals
from
the
3M
Decatur
plant.
Two
lots
of
30
fish
were
used.
One
lot
was
exposed
to
Decatur
plant
effluent
for
21
days
and
the
other
to
river
water
only
for
23
days.
Exposed
fish,
both
living
and
dead,
as
well
as
the
control
fish
were
homogenized
and
analyzed
for
fluorochemicals
by
GC,
TLC,
and
GC/
MS.
There
were
no
detectable
amounts
of
APFO
in
the
ethyl
acetate
or
toluene
extracts
of
the
tissues.
No
fluorochemicals
were
detected
in
the
river
water
exposed
fish.
However,
interpretation
of
this
study
is
problematic
for
several
reasons.
Effluent
concentrations
of
subject
fluorochemicals
were
not
characterized
and
the
specific
protocol
for
exposure
of
the
fish
was
not
found.
There
was
also
no
information
on
analysis
of
the
Tennessee
River
water
or
effluent
used
in
the
study.
Additionally,
it
was
not
known
if
there
was
any
opportunity
for
the
depuration
of
the
fish
prior
to
sacrifice.
No
explanation
was
attempted
as
to
what
was
the
cause
15
of
the
twelve
dead
fish
in
the
effluent­
exposed
group.
The
study
also
did
not
differentiate
between
the
bioaccumulation
of
the
test
compound
and
the
sorption
onto
the
surface
of
the
fish.

2.2.6
Soil
Adsorption
The
adsorption­
desorption
of
APFO
was
studied
in
25
ml
solutions
of
14C­
labeled
APFO
in
distilled
water
with
5
g
Brill
sandy
loam
soil
for
24
hours
at
a
temperature
of
16­
19
º
C.
The
study
reported
a
Kd
of
0.21
and
a
Koc
of
14
indicating
that
PFOA
has
high
mobility
in
Brill
sandy
loam
soil
(
Welsh
1978).
The
Koc
value,
however,
is
questionable
due
to
the
lack
of
accurate
information
on
the
purity
of
the
14C­
labeled
test
substance
(
Boyd
1993a,
b).

Moody
and
Field
(
1999)
conducted
sampling
and
analysis
of
samples
taken
from
groundwater
1
to
3
meters
below
the
soil
surface
in
close
proximity
to
two
fire­
training
areas
with
a
history
of
aqueous
film
forming
foam
use.
Perfluorooctanoate
was
detected
at
maximum
concentrations
ranging
from
116
to
6750
ug/
L
at
the
two
sites
many
years
after
its
use
at
those
sites
had
been
discontinued.
These
results
suggest
that
APFO
may
have
the
potential
to
migrate
through
soils
to
relatively
shallow
groundwater
where
it
persists.

2.3
Environmental
Exposure
2.3.1
Discharge
to
Air
For
1997,
3M
estimated
1950
lbs.
of
PFOA­
compound
(
PFOA
and
related
salts)
stack
releases
at
its
Cottage
Grove,
MN
location
and
another
4500
lbs.
from
Cottage
Grove
incinerated
offsite
(
3M,
2000).
In
1998,
70%
of
the
fluoride­
containing
wastes
at
3M's
Decatur
location
were
incinerated
off­
site;
incineration
is
now
the
primary
disposal
method
for
these
materials
(
3M,
2000).
For
1999,
DuPont
estimated
stack
releases
of
24,000
lbs.
APFO
at
its
Washington
Works,
WV
location,
plus
another
16,000
lbs.
from
Washington
Works
incinerated
offsite
(
DuPont,
2000).

The
above
data
may
not
account
for
all
of
the
sources
of
air
releases
of
PFOA.
Other
potential
sources
addressed
in
the
literature
include
thermolysis
of
fluoropolymers
(
Ellis
et
al,
2001);
the
low
temperature
of
sublimation
(~
135C)
that
would
yield
PFOA
from
drying,
aka
sintering,
of
fluoropolymer
made
with
PFOA
as
a
processing
aid;
the
thermally
refractive
character
of
the
CF
bond
(
Napoli,
et
al,
1984;
Taylor,
et
al,
1990;
Tsang,
et
al,
1998)
that
would
require
very
high
temperature
incineration
as
a
means
of
destruction
of
this
molecule.

2.3.2
Discharge
to
Water
By
analogy
to
PFOS,
PFOA
discharged
to
water
may
remain
there,
become
adsorbed
to
particulate
matter
and
sediment,
and/
or
be
assimilated
by
organisms.
For
1999,
3M
estimated
PFOA­
compound
water
releases
of
<
30,000
lbs.
at
its
Decatur
AL
location,
and
<
15,000
lbs.
at
its
Cottage
Grove
MN
location
(
3M
Company,
2000a,
b).
For
1999,
DuPont
estimated
the
following
APFO
water
releases
per
location:
Washington
Works
WV,
55,000
lbs;
Parlin
NJ,
300
16
lbs.;
Spruance
VA,
150
lbs.;
Chambers
Works
NJ,
9500
lbs.
(
DuPont,
2000).

DuPont
measured
an
APFO
concentration
at
its
site
Washington
Works,
WV
site
of
0.552
ug/
l
from
a
1999
drinking
water
sample
obtained
from
GE
Plastics
immediately
downstream
on
the
Ohio
River.
Modeled
1996
APFO­
compound
releases
indicated
an
average
annual
PFOA
concentration
of
0.423
ug/
l,
with
APFO
concentrations
likely
to
exceed
1
ug
C­
8/
l
about
50%
of
the
time
during
the
year,
and
likely
to
exceed
10
ug
APFO/
L
about
2.2%
of
the
time
during
the
year.

2.3.3
Discharge
to
Land
3M
reported
that
land
treatment
of
sludge
from
wastewater
treatment
at
their
Decatur,
AL
location
ended
in
mid­
1998;
less
than
500
lbs.
were
disposed
to
land
at
that
site
in
1997.
Sludge
from
the
Decatur
site
is
now
transported
to
an
offsite
landfill;
sludge
from
3M's
Cottage
Grove,
MN
facility
is
sent
to
an
industrial
landfill
(
3M
Company,
2000a,
b).
DuPont
(
2000)
estimated
3,900
lbs.
of
APFO
sludge
landfilled
on
site
in
1999
at
their
Chambers
Works,
NJ
facility.
DuPont
estimated
2,600
lbs.
APFO
transferred
offsite
to
a
hazardous
waste
landfill
from
their
Washington
Works,
WV
facility.

Prior
operations
resulted
in
ground­
and
surface
water
concentrations
of
APFO
monitored
at
three
landfills
operated
by
DuPont's
Washington
Works
WV
facility.
Average
surface
water
concentrations
for
two
landfills
were
1392
ug/
L
and
18.5
ug/
L,
respectively.
A
third
landfill
had
a
maximum
concentration
of
33
ug/
L
in
the
permitted
outfall.
Average
groundwater
concentrations
for
two
landfills
were
2537
ug/
L
and
8.83
ug/
L,
respectively.
A
third
landfill
had
a
maximum
groundwater
concentration
of
15
ug/
L
(
DuPont,
2000).

DuPont
also
reported
the
following
APFO
concentrations,
measured
January
2000,
in
three
drinking
water
wells
of
the
Lubeck
Public
Service
District,
downstream
of
DuPont's
Washington
Works
WV
site:
0.8
ug/
L,
0.44
ug/
L,
and
0.313
ug/
L
(
DuPont,
2000).
As
of
August
2000,
the
Lubeck
Public
Service
District
(
LPSD)
reported
APFO
concentrations
of
0.2
ppb
in
drinking
water
at
DuPont's
Washington
Works
facility,
and
0.2,
0.5,
and
0.1
ppb
in
the
three
LPSD
wells
(
LPSD,
2000).

2.3.4
Environmental
Monitoring
3M's
Multi­
City
Study
reported
on
PFOA
concentrations
from
water,
sludge,
sediment,
POTW
effluent,
and
landfill
leachate
samples
taken
in
six
cities
(
3M,
2001a).
Four
of
the
cities
(
Decatur,
AL;
Mobile,
AL;
Columbus,
GA;
Pensacola,
FL)
were
"
supply"
cities
that
have
manufacturing
or
industrial
use
of
fluorochemicals;
two
of
the
cities
(
Cleveland,
TN;
Port
St.
Lucie,
FL)
were
"
control"
cities
that
do
not
have
significant
fluorochemical
activities.
Across
all
cities,
POTW
effluent
concentrations
ranged
from
0.040
to
2.42
ppb.
The
POTW
sludge
(
dry
wt.)
range
was
non­
detect
to
244
ppb;
the
drinking
water
range
was
non­
detect
to
0.029
ppb;
the
landfill
leachate
range
was
non­
detect
to
48.1
ppb;
the
surface
water
range
was
non­
detect
to
0.083;
the
sediment
range
was
non­
detect
to
1.75
ppb
(
dry
wt.);
and
the
quiet
water
range
was
17
non­
detect
to
0.097
ppb.
The
"
control"
cities
samples
generally
inhabited
the
lower
end
of
the
above
ranges,
except
for
the
POTW
effluent
and
sludge
findings
for
Cleveland,
which
were
intermediate
in
their
ranges.

The
Multi­
City
Study
also
included
a
market
basket
sampling
of
PFOA
residue
in
a
total
of
over
200
samples
taken
from
green
beans,
apples,
pork
muscle,
cow
milk,
chicken
muscle,
chicken
eggs,
bread,
hot
dogs,
catfish,
and
ground
beef
(
3M,
2001a).
Measurable
quantities
of
PFOA,
ranging
up
to
2.35
ng/
g,
were
found
in
two
ground
beef
samples
(
control
cities),
two
bread
samples
(
control
and
supply
cities),
two
apple
samples
(
supply
cities),
and
one
green
bean
sample
(
supply
city).

Giesy
reported
that
PFOA
was
rarely
found
in
fish
and
fish­
eating
water
birds.
Fish
were
sampled
from
the
U.
S.,
certain
European
countries,
the
North
Pacific
Ocean,
and
Antarctic
locations
(
Giesy,
2001a).
Fish­
eating
bird
samples
were
collected
from
the
U.
S.,
including
Midway
atoll,
the
Baltic
and
Mediterranean
Seas,
Japanese
and
Korean
coasts
(
Giesy,
2001b)

Giesy
reported
on
PFOA
in
mink
and
river
otter
livers
from
the
U.
S.
(
Giesy,
2001c).
PFOA
was
found
in
a
few
mink
livers
from
Massachusetts
at
a
concentration
range
of
<
18
to
108
ng/
g,
dry
wt.,
but
not
found
in
mink
from
Louisiana,
South
Carolina
and
Illinois.
PFOA
concentrations
in
river
otter
livers
from
Washington
and
Oregon
States
were
less
than
the
quantification
limit
of
36
ng/
g,
wet
wt.

Giesy
reported
that
PFOA
was
not
detected
at
quantifiable
concentrations
in
oysters
collected
in
the
Chesapeake
Bay
and
Gulf
of
Mexico
of
the
U.
S.
coast
(
Giesy,
2001d).

Giesy
reported
on
the
concentrations
of
PFOA
in
surface
water,
sediments,
clams,
and
fish
collected
from
locations
upstream
and
downstream
of
the
3M
facility
at
Decatur,
AL
(
Giesy,
2001e).
Of
the
three
downstream
water
and
sediment
sampling
locations,
the
two
closest
to
the
3M
facility
had
PFOA
surface
water
concentrations
significantly
greater
than
the
two
upstream
sites
(
means
of
1900ug/
L
and
1024
ug/
L,
vs.
0.008
(
est.)
and
0.028
ug/
L);
the
three
downstream
locations
also
had
sediment
concentrations
significantly
greater
than
the
upstream
sites
(
wet
wt.
means
1855
ug/
kg,
892
ug/
kg,
238
ug/
kg
vs.
0.08(
est.)
and
0.09(
est.)).
The
average
fish
whole
body
PFOA
concentration
for
the
upstream
location
was
11.7
ug/
kg
(
wet
wt.),
while
that
for
the
downstream
location
was
106.4
ug/
kg;
the
small
sample
size
prevented
determination
of
significance.
The
average
PFOA
concentration
in
clams
at
the
upstream
location
was
4.38
ug/
kg;
that
for
the
downstream
location
was
8.42
ug/
kg.
These
differences
were
not
significant.

Hansen
(
2002)
reported
concentrations
of
PFOA
measured
from
surface
water
samples
taken
from
the
Tennessee
River
up­
and
downstream
of
the
outfall
from
the
fluorochemical
manufacturing
facility
at
Decatur
AL
(
the
3M
facility
mentioned
above).
There
were
20
sampling
sites
above
and
20
sites
below
the
outfall
location,
spaced
at
approximately
2
mile
intervals.
None
of
the
samples
taken
from
upstream
of
the
facility
had
measurable
concentrations
of
PFOA.
The
downstream
concentrations
of
PFOA
were
observed
to
increase
at
a
point
approximately
six
miles
below
the
outfall;
the
average
PFOA
concentration
from
that
18
point
downstream
was
394
+
128
ng/
L.
The
report
states
that
the
consistency
of
the
PFOA
concentrations
within
these
two
regions
suggests
the
absence
of
either
major
environmental
sinks
or
additional
sources
of
PFOA
in
the
areas
sampled.

Moody
reported
the
concentrations
of
PFOA
in
surface
water
from
stream
locations
sampled
upstream
and
downstream
of
a
spill
of
fire­
fighting
foam
that
contained
perfluorinated
surfactants,
including
PFOA
(
Moody
et
al.,
2002).
Upstream
surface
water
samples
taken
over
the
three
week
period
post­
spill
had
PFOA
levels
of
0.008­
0.033
ug/
L;
corresponding
downstream
levels
ranged
from
0.035­
10.6
ug/
L.
Statistical
significance
was
not
evaluated
for
these
measurements.

2.4
Human
Biomonitoring
Table
2
provides
serum
PFOA
levels
in
both
occupational
cohorts
and
in
the
general
population.
The
highest
levels
reported
to
date
in
the
general
population
are
similar
to
some
of
the
lowest
levels
in
workers
exposed
to
PFOA
occupationally.
The
data
are
currently
limited
to
those
discussed
below.

3M
has
been
offering
voluntary
medical
surveillance
to
workers
at
plants
that
produce
or
use
perfluorinated
compounds
since
1976.
Serum
PFOA
levels
have
been
measured
and
reported
since
1993.
Prior
to
this
time,
only
total
organic
fluorine
was
measured.
The
results
of
biomonitoring
for
PFOA
have
been
reported
for
3
plants:
Cottage
Grove,
MN;
Decatur,
AL
and
Antwerp,
Belgium.
Surveillance
years
include
1993,
1995,
1997,
1998,
and
2000,
although
not
all
of
the
plants
offered
surveillance
in
all
of
these
years.
The
1998
data
reported
for
the
Decatur
plant
consist
of
a
random
sample
of
employees;
however,
volunteers
participated
in
all
of
the
other
sampling
periods
for
all
of
the
plants.

Mean
serum
PFOA
levels
have
increased
slightly
at
both
the
Cottage
Grove
and
Decatur
plants
since
1993.
Workers
at
the
Cottage
Grove
plant,
where
PFOA
exposures
are
highest,
have
the
highest
PFOA
serum
levels.
The
latest
sample
was
in
1997
(
Olsen,
et
al.,
1998b).
The
mean
serum
PFOA
level
was
6.4
ppm
(
range
=
0.1
 
81.3
ppm).
Only
74
employees
participated
in
the
1997
surveillance.
The
eligible
voluntary
participation
rates
ranged
from
approximately
50%
in
1997
to
70%
in
1993.

At
the
Decatur
plant,
263
of
500
employees
participated
in
2000
(
Olsen,
et
al.,
2001a).
The
mean
serum
PFOA
level
was
1.78
ppm.
It
was
higher
in
males
(
n
=
215)
than
females
(
n
=
48),
1.90
and
1.23
ppm,
respectively.
In
addition,
male
production
employees
had
higher
mean
serum
levels
(
2.34
ppm).
Five
employees
had
serum
levels
greater
than
5
ppm,
the
Biological
Limit
Value
established
by
the
3M
Exposure
Guideline
Committee.
Cell
operators
had
the
largest
increase
in
serum
PFOA
between
1998
and
2000.
The
highest
level
was
in
a
chemical
operator
on
the
Scotchgard
team
(
12.70
ppm).
The
mean
level
for
the
rest
of
the
members
of
the
team
was
5.06
ppm
(
range
5
­
9
ppm).
Other
job
categories
did
not
exhibit
such
a
large
increase.
3M
reports
that
this
is
due
to
increased
PFOA
production
at
the
Decatur
plant
beginning
in
1999.
19
Serum
PFOA
levels
for
the
Antwerp
plant
are
lower
than
at
Decatur
or
Cottage
Grove,
and
have
decreased
slightly
since
1995
(
Olsen,
et
al.,
2001b).
Participation
in
medical
surveillance
at
the
Antwerp
plant
was
the
highest
it
had
ever
been
in
2000
(
258
volunteers
out
of
340
workers).
The
mean
serum
PFOA
level
was
0.84
ppm,
and
the
highest
serum
level
reported
was
7.04
ppm.
As
in
the
Decatur
plant,
males
(
n
=
209)
had
higher
mean
serum
PFOA
levels
(
1.03
ppm)
than
females
(
n
=
49,
0.07
ppm).
Three
employees
had
levels
greater
than
5
ppm.

3M's
Specialty
Materials
Manufacturing
Division
Laboratories,
where
employees
perform
fluorochemical
research
(
Building
236),
conducted
voluntary
biomonitoring
of
45
employees
in
2000
(
Olsen,
et
al.,
2001c).
The
mean
PFOA
serum
level
was
0.106
ppm
(
range
0.008
 
0.668
ppm).

Serum
PFOA
levels
in
corporate
staff
and
managers
at
a
3M
plant
in
St.
Paul,
MN,
where
occupational
exposure
to
PFOA
should
not
have
occurred,
were
reported
(
3M
Report,
1999).
Four
of
31
employees
had
serum
PFOA
levels
greater
than
the
detection
limit
of
10
ppb.
The
mean
for
these
employees
was
12.5
ppb.

Data
on
PFOA
levels
in
the
general
population
are
very
limited.
They
are
very
recent
so
that
trends
over
time
cannot
be
established.
The
mean
serum
PFOA
levels
are
lower
in
the
general
population
than
in
workers
exposed
to
PFOA.

Pooled
blood
samples
from
U.
S.
blood
banks
indicate
mean
PFOA
levels
of
3
to
17
ppb
(
3M
Company,
Feb.
5,
1999;
3M
Company,
May
26,
1999).
The
highest
pooled
sample
reported
was
22
ppb.
Samples
were
collected
in
1998
and
1999.
However,
it
cannot
be
assumed
that
these
levels
are
generalizable
to
the
U.
S.
population
for
several
reasons:
1)
blood
donors
are
a
unique
group
that
does
not
necessarily
reflect
the
U.
S.
population
as
a
whole,
2)
many
of
the
blood
banks
originally
contacted
for
possible
inclusion
in
the
study
declined
to
participate,
3)
only
a
small
number
of
samples
have
actually
been
analyzed
for
PFOA,
and
4)
no
other
data
such
as
age,
sex,
or
other
demographic
information
are
available
on
the
donors.

Individual
blood
samples
from
3
different
age
populations
were
recently
analyzed
for
PFOA
and
other
fluorochemicals
using
high­
pressure
liquid
chromatography/
electrospray
tandem
mass
spectrometry
(
HPLC/
ESMSMS)
(
Olsen,
et
al.,
2002a,
2002b,
2002c).
The
studies'
participants
included
adult
blood
donors,
an
elderly
population
participating
in
a
prospective
study
in
Seattle,
WA,
and
children
from
23
states
participating
in
a
clinical
trial.
Overall,
the
PFOA
geometric
means
were
similar
across
all
3
populations
(
4.6
ppb,
4.2
ppb,
and
4.9
ppb,
respectively).
The
geometric
means
and
95%
tolerance
limits
(
the
proportion
of
the
population
expected
to
be
found)
and
their
upper
bounds
were
comparable
across
all
3
studies.
However,
the
upper
ranges
for
the
children
and
adults
were
much
higher
than
for
the
elderly
population.
It
is
not
clear
whether
this
is
the
result
of
geographic
differences
in
PFOA
levels
or
some
other
factor.
It
should
be
noted
that
PFOS
and
PFOA
were
highly
correlated
in
all
three
studies
(
r
=
.63,
r
=
.70,
and
r
=
.75)
and
that
PFOA
did
not
meet
the
criteria
for
a
log
normal
distribution
based
on
the
Shapiro­
Wilk
test
in
any
of
the
studies.
The
authors
suggest
that
it
may
be
due
to
the
greater
20
proportion
of
subjects
with
values
<
LLOQ;
however,
only
geometric
means
were
reported.
The
details
of
each
study
are
provided
below.

Blood
samples
from
645
U.
S.
adult
blood
donors
(
332
males,
313
females),
ages
20­
69,
were
obtained
from
six
American
Red
Cross
blood
banks
located
in:
Los
Angeles,
CA;
Minneapolis/
St.
Paul,
MN;
Charlotte,
NC;
Boston,
MA;
Portland,
OR,
and
Hagerstown,
MD
(
Olsen,
et
al.,
2002a).
Each
blood
bank
was
requested
to
provide
approximately
10
samples
per
10­
year
age
intervals
(
20­
29,
30­
39,
etc.)
for
each
sex.
The
only
demographic
factors
known
for
each
donor
were
age,
gender,
and
location.

The
geometric
mean
serum
PFOA
level
was
4.6
ppb.
The
range
was
<
lower
limit
of
quantitation
(
1.9)
to
52.3
ppb.
Males
had
significantly
higher
(
p
<
.05)
geometric
mean
PFOA
levels
than
females
(
37.8
ppb
vs.
32.1
ppb).
Age
was
not
an
important
predictor
of
adult
serum
fluorochemical
concentrations.
When
stratified
by
geographic
location,
the
highest
geometric
mean
for
PFOA
was
in
the
samples
from
Charlotte,
NC
(
6.3
ppb,
range:
2.1
 
29.0)
and
the
lowest
from
Portland
(
3.6
ppb,
range:
2.1
 
16.7).
The
highest
individual
value
was
reported
in
Hagerstown
(
52.3
ppb).

Serum
PFOA
levels
were
reported
for
238
(
118
males
and
120
females)
elderly
volunteers
in
Seattle
participating
in
a
study
designed
to
examine
cognitive
function
in
adults
aged
65­
96
(
Olsen,
et
al.,
2002b).
Age,
gender
and
number
of
years'
residence
in
Seattle
were
the
only
data
available
on
the
participants.
Most
of
the
participants
were
under
the
age
of
85
and
had
lived
in
the
Seattle
area
for
over
50
years.

The
geometric
mean
of
PFOA
for
all
samples
was
4.2
ppb
(
95%
CI,
3.9
 
4.5).
The
range
was
1.4
 
16.7
ppb.
There
was
no
significant
(
p
<
.05)
difference
in
geometric
means
for
males
and
females.
In
simple
linear
regression
analyses,
age
was
negatively
(
p
<
.05)
associated
with
PFOA
in
elderly
men
and
women.
In
bootstrap
analyses,
the
mean
of
the
95%
tolerance
limit
for
PFOA
was
9.7
ppb
with
an
upper
95%
confidence
limit
of
11.3
ppb.
PFOS
and
PFOA
were
highly
correlated
(
r
=
.75)
in
this
study.

A
sample
of
598
children,
ages
2­
12
years
old,
participating
in
a
study
of
group
A
streptococcal
infections,
was
analyzed
for
serum
PFOA
levels
(
Olsen,
et
al.,
2002c).
The
samples
were
collected
in
1994­
1995
from
children
residing
in
23
states
and
the
District
of
Columbia.
PFOA
did
not
meet
the
criteria
for
a
log
normal
distribution
based
on
the
Shapiro­
Wilk
test.
The
authors
suggest
that
it
may
be
due
to
the
greater
proportion
of
subjects
with
values
<
LLOQ
for
PFOA.
However,
only
geometric
means
were
reported.
The
geometric
mean
of
PFOA
for
all
of
the
participants
was
4.9
ppb
(
95%
CI,
4.7
 
5.1).
The
range
was
1.9
to
56.1
ppb.
Male
children
had
significantly
(
p<.
01)
higher
geometric
mean
serum
PFOA
levels
than
females:
5.2
ppb
and
4.7
ppb,
respectively.
In
simple
linear
regression
analyses,
age
was
significantly
(
p
<
.05)
negatively
associated
with
PFOA
in
both
males
and
females.
When
stratified
by
age,
the
geometric
mean
of
PFOA
was
highest
at
age
4
(
5.7
ppb)
and
lowest
at
age
12
(
3.5
ppb).
Although
the
data
were
not
reported,
a
graphical
presentation
of
log
PFOA
levels
for
each
state
by
gender
looked
similar
across
the
states;
however,
it
is
difficult
to
interpret
these
data
without
21
the
data
and
given
the
limited
sample
size
for
each
gender/
location
subgroup.
In
bootstrap
analyses,
the
mean
of
the
95%
tolerance
limit
for
PFOA
was
10.1
ppb
with
an
upper
95%
confidence
limit
of
11.0
ppb.
PFOS
and
PFOA
were
highly
correlated
(
r
=
.70)
in
this
study.
PFOA
and
PFHS
(
perfluorohexanesulfonate)
were
also
correlated
(
r
=
.48).

The
above
3
studies
indicate
similar
geometric
means
and
ranges
of
PFOA
among
sampled
adults,
children,
and
an
elderly
population.
However,
an
unexpected
finding
was
the
level
of
PFHS
and
M570
(
N­
methyl
perfluorooctanesulfonamidoacetate)
in
children.
These
serum
levels
were
much
higher
in
the
sampled
children
than
in
the
sampled
adults
or
elderly.
It
is
not
clear
why
this
occurred,
but
it
is
probably
due
to
a
different
exposure
pattern
in
children.

In
another
study,
the
PFOA
concentration
was
analyzed
in
human
sera
and
liver
samples
(
Olsen
et
al.,
2001d).
Thirty­
one
donor
samples
were
obtained
from
16
males
and
15
females
over
an
18­
month
period
from
the
International
Institute
for
the
Advancement
of
Medicine
(
IIAM).
The
average
age
of
the
male
donors
was
50
years
(
SD
15.6,
range
5­
69)
and
the
average
age
of
the
female
donors
was
45
years
(
SD
18.5,
range
13­
74).
The
causes
of
death
were
intracranial
hemorrhage
(
n
=
16
or
52%),
motor
vehicle
accident
(
n
=
7
or
23%),
head
trauma
(
n
=
4
or
13%),
brain
tumor
(
n
=
2
or
6%),
drug
overdose
(
n
=
1
or
3%)
and
respiratory
arrest
(
n
=
1
or
3%).
Both
serum
and
liver
tissue
were
obtained
from
23
donors;
7
donors
contributed
liver
tissue
only
and
1
donor
contributed
serum
only.
Serum
samples
were
obtained
from
5
ml
of
blood;
liver
samples
consisted
of
10
g
of
tissue.
Samples
were
frozen
at
IIAM
and
shipped
frozen
to
3M
for
analysis.
Samples
were
extracted
using
an
ion­
pairing
extraction
procedure
and
were
quantitatively
assayed
using
HPLC­
ESMSMS
and
evaluated
versus
an
unextracted
curve.
Extensive
matrix
spike
studies
were
performed
to
evaluate
the
precision
and
accuracy
of
the
extraction
procedure.
Serum
values
for
PFOA
ranged
from
<
LOQ
(<
3.0)
 
7.0
ng/
mL.
Assuming
the
midpoint
value
between
zero
and
LOQ
serum
value
for
samples
<
LOQ,
the
mean
serum
PFOA
level
was
3.1
ng/
mL
with
a
geometric
mean
of
2.5
ng/
mL.
No
liver
to
serum
ratios
were
provided
because
more
than
90%
of
the
individual
liver
samples
were
<
LOQ.
22
Table
2.
Serum
PFOA
Levels
in
Human
Populations
*
Geometric
mean
and
95%
confidence
intervals
were
not
included
in
the
reports.
**
PFOA
detected
in
about
1/
3
of
the
pooled
samples
but
quantifiable
in
only
2
***
only
4
employees
were
above
the
detection
limit
of
10
ppb
Occupational
Exposures
 
Serum
Levels
(
ppm)

Plant
Arithmetic
Mean
Range
Geometric
Mean
95%
CI
Cottage
Grove
Plant
1997
(
n
=
74)
1995
(
n
=
80)
1993
(
n
=
111)
6.4
6.8
5.0
0.1
 
81.3
0.0
 
114.1
0.0
 
80.0
*
*
*
*
*
*
Decatur
Plant
2000
(
n
=
263)
1998
(
n
=
126)
1997
(
n
=
84)
1995
(
n
=
90)
1.78
1.54
1.57
1.46
0.04
 
12.70
0.02
 
6.76
not
reported
not
reported
1.13
0.90
*
*
0.99
 
1.30
0.72
 
1.12
*
*
Antwerp
Plant
2000
(
n
=
258)
1995
(
n
=
93)
0.84
1.13
0.01
 
7.04
0.00
 
13.2
0.33
*
0.27
 
0.40
*
Building
236
2000
(
n
=
45)
0.106
0.008
 
0.668
0.053
0.037
 
0.076
General
Population
Exposures
 
Serum
Levels
(
ppb)
Source
Arithmetic
Mean
Range
Geometric
Mean
95%
CI
Pooled
samples
Commercial
sources
of
blood,
1999
(
n
=
35
lots)
3
1
­
13
*
*

Blood
Banks
(
n
=
18),
1998
~
340­
680
donors
17**
12
­
22
*
*
Individual
samples
American
Red
Cross
blood
banks,
2000
(
n
=
645)
5.6
1.9
­
52.3
4.6
4.3
 
4.8
Elderly
(
ages
65­
96),
2000
(
n
=
238)
not
reported
1.4
 
16.7
4.2
3.9
 
4.5
Children
(
ages
2­
12),
1995
(
n=
598)
5.6
1.9
­
56.1
4.9
4.7
 
5.1
3M
Corporate
managers/
staff
St.
Paul,
MN,
1998
(
n
=
31)
12.5***
not
reported
*
*
23
3.0
Human
Health
Hazards
3.1.
Metabolism
and
Pharmacokinetics
3.1.1
Half­
life
in
Humans
There
are
very
limited
data
on
the
half­
life
of
PFOA.
With
the
exception
of
a
1980
study
in
which
total
organic
fluorine
in
blood
serum
was
measured
in
one
worker,
no
other
data
were
available
until
June
2000
(
Ubel
et
al.,
1980).
A
half­
life
study
on
27
retirees
from
the
Decatur
and
Cottage
Grove
3M
plants
was
undertaken,
in
which
serum
samples
were
drawn
every
6
months
over
a
5­
year
period.
Two
interim
reports
describing
the
results
thus
far
have
been
submitted
(
Burris
et
al.,
2000;
Burris
et
al.,
2002).
The
first
interim
report
suggested
a
median
serum
half­
life
of
PFOA
of
344
days,
with
a
range
of
109
to
1308
days.
The
two
highest
halflife
calculations
were
for
the
2
female
retirees
who
participated
in
this
study
(
654
and
1308
days).

There
were
several
limitations
to
this
first
analysis
including:
1)
the
limited
data
available
and
the
range
of
serum
PFOA
levels
measured;
2)
serum
was
analyzed
after
each
collection
period
with
only
one
measurement
per
time
period
on
different
days
using
slightly
different
analytical
techniques;
and
3)
the
reference
material
purity
was
not
determined
until
after
the
first
3
samples
had
been
analyzed.
An
effort
was
made
to
minimize
experimental
error,
including
systematic
and
random
error
in
the
analytical
method,
involving
9
of
the
original
27
subjects.
Serum
samples
were
collected
from
each
of
the
subjects
over
4
time
periods
spanning
180
days,
measured
in
triplicate
with
all
time
points
from
each
subject
analyzed
in
the
same
analytical
run.
This
would
allow
for
statistical
evaluation
of
the
precision
of
the
measurement
and
assure
that
all
systematic
error
inherent
in
the
assay
equally
affected
each
sample
used
for
half­
life
determination.

Of
the
9
retirees
included
in
this
analysis,
there
were
7
males
and
2
females,
all
from
the
Decatur
plant.
The
average
age
of
the
retirees
was
61
years,
the
mean
number
of
years
worked
at
Decatur
was
27.7
years,
and
the
average
number
of
months
retired
was
18.9.
Average
BMI
of
this
group
was
27.9.
The
mean
PFOA
value
at
study
initiation
was
0.72
ppm
(
range
0.06
 
1.84
ppm,
SD
=
0.64).

The
mean
serum
half­
life
for
PFOA
was
4.37
years
(
range
1.50
 
13.49
years,
SD
=
3.53).
Only
1
employee
had
a
half­
life
value
that
exceeded
4.3
years.
The
2
females
had
values
of
3.1
and
3.9
years.
Age,
BMI,
number
of
years
worked
or
years
since
retirement
were
not
significant
predictors
of
serum
half­
lives
in
multivariable
regression
analyses.

This
analysis
has
attempted
to
reduce
experimental
error
in
the
determination
of
a
half­
life
for
PFOA.
However,
several
issues
should
be
noted.
First,
the
effect
of
continued
nonoccupational
low­
level
exposure
on
the
half­
life
is
unknown.
Second,
because
subjects'
blood
contained
concentrations
of
fluorochemicals
that
varied
by
a
factor
of
30,
the
data
cannot
be
pooled
or
averaged
unless
the
serum
concentration
decay
curve
shows
first­
order
kinetics.
24
Third,
it
is
not
known
if
there
are
interactions
between
PFOA
and
other
fluorochemicals
in
the
body.
Fourth,
this
estimate
is
much
higher
than
that
reported
in
lab
animals.
Fifth,
systematic
error
of
the
analytical
method
could
be
as
high
as
+/­
20%
and
still
satisfy
the
data
quality
criteria.

3.1.2
Absorption
Studies
in
Animals
APFO
is
well
absorbed
following
oral
and
inhalation
exposure,
and
to
a
lesser
extent
following
dermal
exposure.
The
liver
and
serum
levels
have
been
measured
in
several
subchronic
and
reproductive
toxicity
studies.
These
results
are
presented
in
the
relevant
studies
summarized
in
subsequent
sections.
Other
studies
are
summarized
here.

In
female
rats,
an
average
of
749
ug
or
37%
of
the
fluorine
in
the
administered
dose
was
recovered
in
the
urine
within
4.5
hr
after
PFOA
dose
(
by
stomach
intubation
2
ml
of
an
aqueous
solution
containing
2
mg
PFOA)
(
Ophaug
and
Singer,
1980).
The
quantity
of
nonionic
fluorine
recovered
in
the
urine
increased
to
61%
of
the
dose
at
8
hr,
76%
at
24
hr,
and
89%
at
96
hr.

After
a
single
oral
dose
of
14C­
PFOA
(
mean
dose,
11.0
mg/
kg)
in
solution
to
groups
of
three
male
rats,
at
least
93%
of
the
total
carbon­
14
was
absorbed
at
24
hours
(
Gibson
and
Johnson,
1979).
The
half­
life
for
elimination
of
total
carbon­
14
from
plasma
was
4.8
days.

Following
APFO
head­
only
inhalation
exposure
in
male
rats
(
6
hr/
day,
5
days/
wk
for
2
wk
to
0,
1,
8
or
84
mg/
m3)
concentrations
of
organofluoride
in
the
blood
showed
a
dose
relationship
with
initial
levels
of
108
ppm
in
rats
treated
at
84
mg/
m3
(
Kennedy
et
al.,
1986).
Immediately
after
the
tenth
exposure
period,
the
mean
organofluoride
blood
levels
were
13
ppm,
47
ppm,
and
108
ppm
in
the
1,
8,
and
84
mg/
m3
dose
groups.

Subchronic
dermal
APFO
treatment
in
rats
and
rabbits
(
10
applications,
5
doses,
2
rest
days,
5
doses)
with
either
0,
20,
200,
or
2000
mg/
kg
resulted
in
elevated
blood
organofluorine
levels
which
increased
in
a
dose­
related
manner
(
Kennedy,
1985).

O'Malley
and
Ebbins
(
1981)
conducted
a
range
finding
study
which
indicates
significant
dermal
absorption
of
PFOA
in
male
and
female
rabbits.
PFOA
(
100
mg/
kg,
1000
mg/
kg,
and
2000
mg/
kg
in
saline
slurry)
was
applied
to
approximately
40%
of
the
shaved
trunk
of
the
animals,
which
were
then
fitted
with
a
plastic
collar,
and
the
trunk
was
wrapped
with
impervious
plastic
sheeting.
The
exposure
period
was
24
hr,
5
days/
week
over
14
days.
Mortality
was
100%
(
4/
4)
in
the
2000
mg/
kg
group,
75%
(
3/
4)
in
the
1000
mg/
kg
group
and
0%
(
0/
4)
in
the
100
mg/
kg
group.

A
tetrabutyl
ammonium
salt
of
perfluorooctanoate
in
the
form
of
treated
fabric
and
as
a
liquid
formulation
was
applied
dermally
to
rabbits
(
Johnson,
1995b).
Liver
samples
were
analyzed
at
28
days
post
dose
for
total
organic
fluorine.
The
results
from
treated
animals
were
the
same
as
control
values.
All
total
organic
values
were
below
the
practical
quantitation
limit.
Serum
levels
were
also
below
the
practical
quantitation
limits
of
the
analysis
for
samples
25
collected
at
day
1
and
2
after
administration
of
the
mixture
or
the
treated
fabric.
From
the
pharmacokinetic
study
(
Johnson,
1995a),
it
would
be
unlikely
that
any
extent
of
absorption
could
have
been
detected
in
this
study.

3.1.3
Distribution
Studies
in
Animals
PFOA
distributes
primarily
to
the
liver,
plasma,
and
kidney,
and
to
a
lesser
extent,
other
tissues
of
the
body.
It
does
not
partition
to
the
lipid
fraction
or
adipose
tissue,
but
does
bind
to
macromolecules
in
the
tissues.
There
is
evidence
of
enterohepatic
circulation
of
the
compound.

Serum
and
liver
concentrations
of
PFOA
were
determined
in
rhesus
monkeys
in
a
90
day
oral
toxicity
study
(
Griffith
and
Long,
1980).
In
monkeys
at
the
3
mg/
kg/
day
dose,
mean
serum
PFOA
was
50
ppm
in
males
and
58
ppm
in
females.
At
the
same
dose,
males
had
3
ppm
and
females
7
ppm
in
liver
samples.
At
10
mg/
kg/
day
doses,
male
monkeys
had
a
mean
serum
PFOA
of
58
ppm
and
females
75
ppm.
Liver
levels
were
9
and
10
ppm
for
males
and
females,
respectively,
measured
as
organic
fluoride.

Ophaug
and
Singer
(
1980)
measured
ionic
fluoride
and
total
fluorine
in
the
serum
of
female
rats
following
the
administration
of
PFOA
by
stomach
intubation
(
2
ml
of
an
aqueous
solution
containing
2
mg
PFOA).
Serum
from
rats
4.5
hr
after
the
administration
of
PFOA
had
a
nonionic
fluorine
level
13.6
ppm
and
virtually
all
of
this
was
bound
to
components
in
the
serum
and
not
ultrafilterable.
Despite
the
large
increase
in
nonionic
fluorine
in
the
serum,
the
ionic
fluoride
level
remained
very
low
(
0.03
ppm).
Prior
to
intubation
of
PFOA,
the
ionic
and
nonionic
fluorine
levels
in
serum
were
0.032
and
0.07
ppm,
respectively.
The
nonionic
fluorine
level
in
the
serum
decreased
to
11.2
ppm
at
8
hr,
0.35
ppm
at
24
hr,
and
0.08
ppm
at
96
hr.
The
authors
conclude
that
PFOA
is
rapidly
absorbed
from
the
gastrointestinal
tract
and
rapidly
cleared
from
the
serum.

Twenty­
four
hours
after
oral
administration
of
APFO
(
2
mg
APFO
in
2
ml
aqueous
solution
by
stomach
intubation),
female
rats
had
a
mean
serum
nonionic
fluorine
level
of
0.35
ppm,
while
male
rats
had
a
mean
serum
nonionic
fluorine
level
of
44.0
ppm
(
Hanhijarvi
et
al.,
1982).
APFO
was
bound
to
a
similar
extent
in
the
plasma
of
male
and
female
rats
(
97.5%
bound).

In
male
and
female
rats
administered
14C­
PFOA
in
propylene
glycol/
water
(
9.4
umol/
kg,
i.
p.),
the
concentration
of
14C­
PFOA­
derived
radioactivity
in
the
blood
was
higher
and
eliminated
more
slowly
in
males
(
t1/
2=
9
days,
males
vs
4
hr,
females,
Vanden
Heuvel
et
al.,
1991).
In
the
male
rats,
the
liver
had
the
highest
PFOA
concentration
(
21%
of
dose
at
2
hr,
2%
of
dose
at
28
days)
followed
by
the
plasma
and
kidney.
Far
lower
PFOA
concentrations
were
found
in
the
heart,
testis,
fat,
and
gastrocnemius
muscle.
In
females
at
2
hr
post
dose,
the
highest
concentrations
of
PFOA
were
found
in
the
plasma
followed
by
the
kidney,
liver
and
ovaries
in
that
order.
The
average
t1/
2
for
elimination
of
PFOA
from
the
liver
in
male
rats
was
11
days
compared
to
an
average
of
9
days
for
extrahepatic
tissues.
In
females,
the
average
t1/
2
for
tissue
elimination
was
approximately
3
hr.
26
Vanden
Heuvel
et
al.
(
1991)
investigated
the
disposition
of
PFOA
in
perfused
male
rat
liver.
Approximately
11%
of
the
cumulative
dose
of
14C­
PFOA
infused
(
0.08
umol/
min
x
48
min,
3.84
umol
total)
was
extracted
by
the
liver
during
a
first
pass.
In
addition,
the
cumulative
percent
of
PFOA
extracted
by
the
liver
at
2
min
(
33%)
was
substantially
greater
than
that
seen
after
48
min
(
11%)
indicating
that
first­
pass
hepatic
uptake
of
PFOA
may
be
saturable.

Ylinen
et
al.
(
1990)
studied
the
difference
between
male
and
female
Wistar
rats
in
the
distribution
and
accumulation
of
PFOA
after
a
single
and
subchronic
administration.
The
single
dose
of
PFOA
(
50
mg/
kg
in
propylene
glycol­
water
mixture,
1:
1,
vol.
0.25
ml/
100g)
was
administered
intraperitoneally
to
10
week
old
rats
(
20
male,
20
female).
Subchronic
administration
of
PFOA
consisted
of
3,
10,
and
30
mg/
kg/
day
by
gavage
(
in
0.9%
NaCl,
0.5
ml/
100g)
to
newly
weaned
rats
(
18
male,
18
female).
After
the
single
dose,
samples
were
collected
for
PFOA
determination
12,
24­
168
(
at
24
hr
intervals),
244
and
336
hours
after
the
administration,
and
in
the
subchronic
test
on
the
28th
day.
The
serum
was
collected
by
cardiac
puncture;
after
decapitation
the
brain
and
at
necropsy
samples
from
the
liver,
kidney,
lung,
spleen,
ovary,
testis,
and
adipose
tissue
were
collected
and
frozen.
The
biological
half­
life
of
PFOA
in
the
serum
and
tissues
was
determined
from
the
linear
relationship
between
time
and
PFOA
concentration
in
the
semilogarithmic
plot.
In
the
single­
dose
study,
concentration
of
PFOA
in
the
serum
and
tissues
was
higher
in
males
than
females
at
all
time
periods.
Twelve
hours
after
the
administration
of
PFOA
about
10%
of
the
dose
was
found
in
the
serum
of
females,
whereas
about
40%
was
in
the
serum
of
males.
After
14
days
about
3.5%
of
the
dose
remained
in
the
serum.
In
females,
PFOA
concentration
in
the
serum,
liver,
and
kidney
occurred
in
a
discontinuous
fashion,
indicating
distinct
phases.
The
half­
life
in
the
serum
was
24
and
105
h
in
the
females
and
males,
respectively.
In
the
females,
a
half­
life
of
60
hr
was
estimated
in
the
liver
during
the
first
week.
In
the
males,
the
half­
life
in
liver
was
210
hr.
Although
PFOA
was
retained
by
the
liver,
it
was
not
found
in
the
lipid
fraction.
In
the
kidney,
the
half­
life
was
145
hr
and
130
hr
in
females
and
males,
respectively.
In
the
spleen,
the
half­
life
was
73
hr
and
170
hr
in
females
and
males,
respectively.
PFOA
was
also
found
in
brain
tissue.
PFOA
was
not
detectable
in
adipose
tissue.
In
the
subchronic
study,
samples
taken
on
the
28th
day
indicated
significantly
higher
PFOA
concentrations
in
the
serum
and
tissues
of
males
versus
females
in
all
three
dose
levels.
After
subchronic,
as
well
as
single­
dose
administration,
PFOA
was
mainly
distributed
in
the
serum
of
rats.
High
concentrations
of
PFOA
were
also
found
in
the
liver,
kidney,
and
lung
of
males
and
females.
At
the
high
dose
level
(
30
mg/
kg/
day),
females
and
males
exhibited,
respectively,
serum
concentrations
of
13.92
and
51.65
ug/
ml,
liver
concentrations
of
6.64
and
49.77
ug/
g,
kidney
concentrations
of
12.54
and
39.81
ug/
g,
spleen
concentrations
of
1.59
and
4.10
ug/
g,
lung
concentrations
of
0.75
and
23.71
ug/
g,
and
brain
concentrations
of
0.044
and
0.710
ug/
g.
The
ovary
contained
1.16
ug/
g
and
the
testis
contained
7.22
ug/
g.
A
significant
positive
correlation
existed
between
the
administered
dose
and
the
concentration
of
PFOA
in
the
liver,
kidney,
spleen,
and
lung
of
females.
On
the
contrary,
no
significant
correlation
between
the
administered
dose
and
the
concentration
of
PFOA
was
observed
in
the
males,
as
10
mg/
kg/
day
produced
higher
PFOA
concentrations
in
the
serum
and
organs
than
30
mg/
kg/
day.
However,
in
males,
the
concentration
in
the
spleen,
testis,
and
brain
correlated
positively
with
the
concentration
in
the
serum.
27
Vanden
Heuvel
et
al.
(
1992)
demonstrated
that
PFOA
covalently
binds
to
proteins
in
the
liver,
plasma,
and
testes
of
rats.
Carbon­
14­
labeled
PFOA
was
administered
to
six­
week
old
male
Harlan
Sprague­
Dawley
rats
in
propylene
glycol/
water
(
1:
1,
v/
v;
1
ml/
kg)
at
a
dose
of
9.4
umol/
kg,
i.
p.
No
time­
dependent
changes
in
either
absolute
or
relative
concentrations
of
covalently
bound
PFOA­
derived
14C
were
found
at
2
h,
1
and
4
days
post­
treatment.
Covalently
bound
PFOA
was
represented
by
0.1
to
0.3%
of
the
tissue
14C
content.
The
absolute
concentration
of
covalently
bound
PFOA
was
significantly
higher
in
the
plasma
than
in
the
liver.
The
testes
had
the
highest
relative
concentration
of
PFOA­
derived
radioactivity
covalently
bound.
In
tests,
covalent
binding
of
14C­
PFOA
to
a
constant
concentration
of
albumin
(
8
uM)
increased
in
a
linear
fashion
with
increasing
PFOA
concentration.
The
covalent
binding
of
PFOA
to
hemoglobin
was
diminished
by
the
addition
of
cysteine
but
not
methionine,
suggesting
that
protein
sulfhydryl
groups
may
be
involved.

Hanhijarvi
et
al.
(
1987)
compared
the
disposition
of
PFOA
between
male
and
female
Wistar
rats
during
subchronic
administration.
PFOA
was
administered
by
gavage
to
48
newly­
weaned
animals
at
0,
3,
10,
and
30
mg/
kg
(
in
0.9%
NaCl,
0.5ml/
100g)
for
28
consecutive
days.
Urine
was
collected
on
the
7th
and
28th
day
of
the
study
(
discussed
below).
At
the
end
of
the
study,
blood
was
collected
via
cardiac
puncture.
At
each
dose
level,
the
mean
PFOA
concentrations
in
the
plasma
of
the
male
rats
were
significantly
higher
than
those
of
the
female
rats.
The
mean
plasma
PFOA
concentrations
for
the
male
rats
were
48.6+­
26.5
ug/
ml
(
dosed
at
3
mg/
kg),
83.1+­
24.7
ug/
ml
(
10
mg/
kg),
and
53.4+­
11.2
ug/
ml
(
30
mg/
kg).
The
corresponding
figures
for
female
rats
were
2.43+­
5.96
ug/
ml,
11.3+­
8.59
ug/
ml,
and
9.06+­
8.80
ug/
ml
in
the
same
order.
The
PFOA
concentrations
in
the
plasma
of
the
male
animals
suggested
that
the
binding
sites
of
PFOA
may
become
saturated
at
the
chronic
daily
dose
level
of
30
mg/
kg.
Although
the
plasma
PFOA
concentrations
were
significantly
higher
in
the
male
rats,
no
significant
histopathological
differences
between
the
sexes
were
observed
at
necropsy.

The
disposition
of
PFOA
was
studied
in
male
Wistar
rats
after
castration
and
estradiol
administration
as
well
as
in
intact
males
and
females
(
Ylinen
et
al.,
1989).
The
male
rats
(
N=
20)
were
castrated
at
the
age
of
28
days
and
after
5
weeks
were
used
in
the
tests.
Half
of
the
operated
and
10
intact
males
were
administered
estradiol
valerate
subcutaneously
500
ug/
kg
every
second
day
during
14
days
before
the
test.
Blood
samples
were
collected
by
cardiac
puncture.
At
the
end
of
the
test
(
96
hr),
the
concentration
of
PFOA
in
the
serum
of
intact
males
was
considerably
higher
(
17­
40
times)
than
in
the
serum
of
other
groups.
There
was
no
statistically
significant
difference
in
the
serum
concentrations
between
the
other
groups.
PFOA
was
similarly
bound
to
the
proteins
in
the
serum
of
males
and
females.

Johnson
et
al.
(
1984)
investigated
the
effect
of
feeding
cholestyramine
to
rats
on
the
fecal
elimination
of
APFO.
Since
APFO
exists
as
an
anion
at
physiologic
pH,
it
would
be
expected
to
complex
with
cholestyramine.
Ten
male
Charles
River
CD
rats
(
12
weeks
old,
300­
342
g)
were
administered
ammonium
14C­
perfluorooctanoate
(
2.1
mg/
ml)
dissolved
in
0.9%
NaCl
as
a
single
intravenous
dose
(
2
ml/
rat,
average
APFO
dose
13
mg/
kg).
Five
rats
were
given
4%
cholestyramine
in
feed.
Urine
and
feces
samples
were
collected
at
intervals
for
14
days,
at
which
time
the
animals
were
sacrificed
and
liver
samples
were
collected.
At
14
days
post
dose,
28
the
mean
percentage
of
PFOA
dose
eliminated
in
the
feces
of
cholestyramine­
treated
rats
(
43.2+­
5.5)
was
9.8­
fold
the
mean
percentage
of
dose
eliminated
in
feces
by
untreated
rats
(
4.4+­
1.0).
Excretion
in
urine
was
41%
for
treated
rats
and
67%
for
untreated
rats.
Carbon­
14
present
in
the
liver
represented
12.1+­
2.1
ug
eq/
g
and
22.3+­
6.2
ug
eq/
g
in
treated
and
untreated
rats,
respectively
(
4%
and
8%
of
dose,
respectively).
In
plasma,
the
levels
were
5.1+­
1.7
ug
eq/
ml
and
14.7+­
6.8
ug
eq/
ml
in
treated
and
untreated
rats,
respectively.
In
red
blood
cells,
the
levels
were
1.8+­
0.7
ug
eq/
ml
and
4.2+­
2.4
ug
eq/
ml
in
treated
and
untreated
rats,
respectively.
The
high
concentration
of
14C­
APFO
in
liver
at
2
weeks
after
dosing
and
the
fact
that
cholestyramine
treatment
enhances
fecal
elimination
of
carbon­
14
nearly
10­
fold
suggests
that
there
is
enterohepatic
circulation
of
PFOA.

The
disposition
of
the
tetrabutyl
ammonium
salt
of
perfluorooctanoic
acid
in
female
rabbits
has
been
reported
(
Johnson,
1995a).
Individual
rabbits
were
given
intravenous
doses
at
0,
4,
16,
and
24
mg/
kg
and
appeared
normal
throughout
the
study
(
the
animal
treated
at
the
40
mg/
kg
dose
level
died
within
5
minutes
of
dosing).
Serum
samples
were
analyzed
for
total
organic
fluorine
at
2,
4,
6,
8,
12,
24,
and
48
hours
post
dose.
At
2
hrs,
serum
organic
fluorine
levels
in
the
0,
4,
16,
and
24
mg/
kg
dosed
rabbits
were
1.25
ppm,
4.09
ppm,
14.9
ppm,
and
41.0
ppm,
respectively.
There
was
a
rapid
decrease
in
serum
level
of
total
organic
fluorine
with
time,
non­
detectable
at
48
hr.
The
biological
half­
life
was
on
the
order
of
4
hours.
The
total
organic
fluorine
in
whole
liver
at
48
hr
post
dose
for
control
animals,
4
mg/
kg,
16
mg/
kg,
and
24
mg/
kg
intravenous
doses
were
20
ug,
43
ug,
66
ug,
and
54
ug.

3.1.4
Metabolism
Studies
in
Animals
Vanden
Heuvel
et
al.
(
1991)
investigated
the
metabolism
of
PFOA
in
rats
administered
14CPFOA
(
9.4
umol/
kg,
i.
p.).
Pooled
daily
urine
samples
(
0­
4
days
post­
treatment)
and
bile
extracts
analyzed
by
HPLC
contained
a
single
radioactive
peak
eluting
identically
to
the
parent
compound.
Tissues
were
taken
from
rats
treated
4,
14,
and
28
days
previously
with
14C­
PFOA
to
determine
the
presence
of
PFOA­
containing
lipid
conjugates.
Only
the
parent
compound
was
present
in
rat
tissues;
no
PFOA­
containing
hybrid
lipids
were
detected.
Fluoride
concentrations
in
plasma
and
urine
before
and
after
PFOA
treatment
were
unchanged,
indicating
that
PFOA
does
not
undergo
defluorination.

Ophaug
and
Singer
(
1980)
also
found
no
change
in
ionic
fluoride
level
in
the
serum
or
urine
following
oral
administration
of
PFOA
to
female
rats.
Ylinen
et
al.
(
1989)
found
no
evidence
of
phase
II
metabolism
of
PFOA
following
a
single
intraperitoneal
PFOA
dose
(
50
mg/
kg)
in
male
and
female
rats.

3.1.5
Elimination
Studies
in
Animals
There
are
major
gender
differences
in
the
elimination
of
PFOA
in
rats.
The
biological
half­
life
of
PFOA
in
male
rats
is
many
times
greater
than
that
in
female
rats
and
this
difference
is
primarily
due
to
low
renal
clearance
in
male
rats.
The
rapid
excretion
of
PFOA
by
female
rats
is
due
to
active
renal
tubular
secretion
(
organic
acid
transport
system);
this
renal
tubular
secretion
29
is
believed
to
be
hormonally
controlled
since
castrated
male
rats
treated
with
estradiol
have
excretion
rates
of
PFOA
similar
to
those
of
female
rats.
Hormonal
changes
during
pregnancy
do
not
appear
to
change
the
rate
of
elimination
in
rats.
This
gender
difference
has
not
been
observed
in
primates
and
humans.
The
studies
demonstrating
this
are
described
below.

The
urine
is
the
major
route
of
excretion
of
PFOA
in
the
female
rat,
while
the
urine
and
feces
are
both
major
routes
of
excretion
of
PFOA
in
male
rats
(
Vanden
Heuval,
1991).
Male
and
female
rats
were
administered
14C­
PFOA
in
propylene
glycol/
water
(
9.4
umol/
kg,
i.
p.).
Female
rats
eliminated
PFOA­
derived
radioactivity
rapidly
in
the
urine
with
91%
of
the
dose
being
excreted
in
the
first
24
hr,
while
male
rats
excreted
only
6%
of
the
dose
in
that
time
period.
Negligible
radioactivity
was
recovered
in
the
feces
of
female
rats.
In
male
rats
during
the
28­
day
collection
period
the
cumulative
excretion
of
PFOA­
derived
14C
in
urine
and
feces
was
36.4%
and
35.1%,
respectively.
The
female
rat
retained
less
than
10%
of
the
administered
dose
after
24
hr,
while
the
male
rats
retained
30%
of
the
administered
dose
after
28
days.
The
whole­
body
elimination
half­
life
in
females
was
less
than
one
day,
and
in
males
it
was
15
days.
In
renal­
ligated
rats
injected
i.
p.
with
14C­
PFOA,
approximately
0.3%
of
the
PFOA­
derived
radioactivity
was
excreted
in
the
bile
after
6
hr
(
Vanden
Heuvel
et
al.,
1991).
No
sex­
related
difference
in
the
biliary
excretion
of
PFOA
was
observed
when
the
kidneys
were
ligated.

Johnson
and
Gibson
(
1980)
observed
a
sex
difference
in
extent
and
rate
of
excretion
of
total
carbon­
14
between
male
and
female
rats
after
a
single
iv
dose
(
mean
dose:
female,
16.7
mg/
kg;
male
13.1
mg/
kg)
of
14C­
PFOA.
Female
rats
excreted
essentially
all
of
the
dose
via
urine
in
24
hours
while
at
the
same
time
period
male
rats
excreted
only
20
percent
of
the
dose;
male
rats
excreted
83%
via
urine
and
5.4%
via
feces
by
36
days
post
dose.
No
radioactivity
was
detected
in
tissues
of
female
rats
at
17
days
post
dose;
male
rats
had
2.8%
of
the
dose
in
liver
and
1.1%
in
plasma
at
36
days
post
dose
with
lower
levels
(<
0.5%
of
the
dose)
in
other
organs.

Ophaug
and
Singer
(
1980)
investigated
the
metabolic
fate
of
PFOA
in
female
Holtzman
rats.
Animals
weighing
approximately
250
g
were
administered
by
stomach
intubation
2
ml
of
an
aqueous
solution
containing
2
mg
PFOA.
The
animals
were
then
placed
in
metabolism
cages
and
provided
rat
chow
and
tap
water
for
4.5,
8,
24,
or
52.5
hr.
In
addition,
four
rats
were
placed
in
metabolism
cages
and
fed
a
low
fluoride
(<
0.5
ppm)
diet
and
distilled
water
for
a
period
of
96
hr.
At
the
end
of
the
experimental
period
the
urine,
feces
and
serum
were
collected.
Within
4.5
hr
after
PFOA
dose,
an
average
of
749
ug
or
37%
of
the
fluorine
in
the
administered
dose
was
recovered
in
the
urine.
The
quantity
of
nonionic
fluorine
recovered
in
the
urine
increased
to
61%
of
the
dose
at
8
hr,
76%
at
24
hr,
and
89%
at
96
hr.
Urinary
excretion
of
ionic
fluoride
in
the
PFOA
dosed
animals
was
not
significantly
different
than
that
of
the
control
animals.
Fecal
excretion
of
nonionic
fluorine
was
4.5%
of
the
administered
dose
at
52.5
hr
and
14.3%
at
96
hr.
The
urine
from
undosed
animals
contained
no
detectable
nonionic
fluorine.

The
urinary
excretion
of
APFO
in
rats
was
investigated
by
Hanhijarvi
et
al.
(
1982).
Four
male
and
six
female
Holtsman
rats
were
administered
2
mg
APFO
in
2
ml
aqueous
solution
by
stomach
intubation.
Seven
female
rats
were
administered
2
ml
distilled
water
as
controls.
The
animals
were
then
placed
in
metabolism
cages
with
rat
chow
and
tap
water.
Urine
was
collected
30
until
animals
were
sacrificed
at
24
h
by
cardiac
puncture.
Serum
was
collected.
Ionic
fluoride
and
total
fluorine
content
of
serum
and
urine
was
determined,
and
nonionic
fluorine
was
calculated
as
the
difference.
For
clearance
studies
of
APFO
and
inulin,
the
rats
were
anesthetized
with
Inactin.
The
femoral
artery
was
cannulated
for
continuous
infusion
of
5%
mannitol
in
isotonic
saline
and
the
femoral
artery
was
cannulated
for
drawing
blood
samples.
The
urinary
bladder
was
also
cannulated
for
serial
collections
of
urine.
Intravenous
priming
doses
of
5.2­
5.6
mg
[
1­
14C]
ammonium
perfluorooctanoate
(
sp
act
0.5
uCi/
mg)
and
8.8
ug
tritiated
inulin
(
methoxy­
3H,
sp
act
114
uCi/
mg)
were
given
to
each
animal.
The
radiolabled
inulin
and
APFO
in
5%
mannitol
in
isotonic
saline
was
then
infused
at
a
rate
of
0.21
ml/
min.
An
additional
0.42­
0.63
mg/
hr
14C­
APFO
and
9.6
ug/
hr
tritiated
inulin
was
infused
during
the
experiments.
When
the
urine
and
serum
collections
for
the
clearance
study
were
complete,
probenecid
was
administered
(
65­
68
mg/
kg,
ip)
and
additional
clearance
tests
were
performed.
In
the
cumulative
excretion
study,
rats
were
dosed
iv
with
a
mixture
of
radiolabeled
APFO
(
10­
20%)
and
unlabeled
APFO
(
80­
90%).
Five
percent
mannitol
in
isotonic
saline
was
infused
at
a
rate
of
0.081
ml/
min
and
urine
specimens
were
collected
over
30­
min
intervals.
The
effect
of
probenecid
was
assessed
by
administering
65­
68
mg/
kg
ip
at
least
30
min
prior
to
the
administration
of
APFO.
Twenty­
four
hours
after
oral
administration
of
APFO,
female
rats
had
excreted
76+­
2.7%
of
the
dose
in
the
urine
and
had
a
mean
serum
nonionic
fluorine
level
of
0.35+­
0.11
ppm,
while
male
rats
had
excreted
only
9.2+­
3.5%
of
the
dose
and
had
a
mean
serum
nonionic
fluorine
level
of
44.0+­
1.7
ppm.
APFO
was
bound
to
a
similar
extent
in
the
plasma
of
male
and
female
rats
(
97.5+­
0.25%
bound).
The
clearance
studies
demonstrated
major
differences
between
the
sexes
in
rats.
The
APFO
clearance
in
female
rats
was
several
times
greater
than
the
inulin
clearance.
Administration
of
probenecid,
which
strongly
inhibits
the
renal
active
secretion
of
organic
acids,
reduced
APFO/
inulin
clearance
ratio
in
females
from
14.5
to
0.46.
APFO
clearance
was
reduced
from
5.8
to
0.11
ml/
min/
100g.
Net
APFO
excretion
was
reduced
from
4.6
to
0.13
ug/
min/
100g.
In
male
rats,
however,
the
APFO/
inulin
clearance
ratio
and
the
net
excretion
of
APFO
were
virtually
unaffected
by
probenecid.
In
the
males,
APFO
clearance
was
0.17
ml/
min/
100g,
APFO/
inulin
clearance
ratio
was
0.22,
and
net
APFO
excretion
was
0.17
ug/
min/
mg.
In
the
cumulative
excretion
studies,
female
rats
excreted
76%
of
the
APFO
dose,
while
males
excreted
only
7.8%
of
the
dose
over
a
7­
hr
period.
Probenecid
administration
modified
the
cumulative
excretion
curve
for
males
only
slightly.
However,
in
females
probenecid
markedly
reduced
APFO
elimination
to
11.8%.
The
authors
concluded
that
the
female
rat
possesses
an
active
secretory
mechanism
which
rapidly
eliminates
APFO
from
the
body.
This
secretory
mechanism
is
lacking
or
is
relatively
inactive
in
male
rats
and
accounts
for
the
greater
toxicity
of
APFO
in
male
rats.

Hanhijarvi
et
al.
(
1987)
compared
the
urinary
elimination
of
PFOA
between
male
and
female
Wistar
rats
during
subchronic
administration.
APFO
was
administered
by
gavage
to
48
newlyweaned
animals
at
0,
3,
10,
and
30
mg/
kg
(
in
0.9%
NaCl,
0.5ml/
100g)
for
28
consecutive
days.
Urine
was
collected
on
the
7th
and
28th
day
of
the
study.
At
the
end
of
the
study,
blood
was
collected
via
cardiac
puncture.
At
necropsy,
tissue
specimens
for
histopathologic
examination
were
collected
from
the
controls
and
from
the
group
receiving
30
mg/
kg/
day
PFOA.
On
the
seventh
day
of
the
study
period,
the
female
rats
in
the
lowest
dose
group
(
3
mg/
kg/
day)
exhibited
significantly
greater
urinary
PFOA
excretion
than
the
males
(
3.12+­
0.30
vs
1.50+­
0.57
31
mg/
24hr/
kg).
Unlike
the
female
rats,
on
the
7th
day
of
the
study
all
three
groups
of
male
rats
excreted
significantly
less
PFOA
than
their
daily
dose
of
PFOA,
which
suggested
that
the
males
had
not
reached
a
steady
state
by
seven
days.
On
the
28th
day,
the
males
excreted
an
amount
of
PFOA
equal
to
their
daily
dose.

Hanhijarvi
et
al.
(
1988)
investigated
the
excretion
kinetics
of
PFOA
in
the
beagle
dog.
Six
laboratory
bred
beagle
dogs
(
3
male,
3
female)
were
anesthetized
with
methoxyflurane
and
catheters
were
placed
in
both
ureters
after
laparototomy
and
cystotomy.
The
animals
were
given
an
intravenous
dose
of
30
mg/
kg
of
PFOA
followed
by
continuous
infusion
with
5%
mannitol
solution
at
1.7
ml/
min.
Urine
was
collected
at
10
minute
intervals
for
60
min.
A
5
ml
blood
sample
was
collected
in
the
middle
of
each
urine
sampling
period.
Probenicid
(
30
mg/
kg
i.
v.)
was
then
administered,
and
urine
and
blood
samples
were
again
collected
as
before.
Renal
clearance
of
PFOA
was
calculated
for
the
before
and
after
probenecid
injection
periods.
Four
additional
dogs
(
2male,
2
female)
were
given
30
mg/
kg
PFOA
intravenously.
These
dogs
were
kept
in
metabolism
cages,
and
blood
samples
were
collected
intermittently
for
30
days.
Renal
clearance
rate
was
approximately
0.03
ml/
min/
kg.
Probenecid
significantly
reduced
the
PFOA
clearance
in
both
sexes,
indicating
an
active
secretion
mechanism
for
PFOA.
The
plasma
halflife
of
PFOA
was
longer
in
the
male
dogs
(
473
h
and
541
h)
than
in
the
female
dogs
(
202
h
and
305
h).

The
urinary
excretion
of
PFOA
was
studied
in
male
Wistar
rats
after
castration
and
estradiol
administration
as
well
as
in
intact
males
and
females
(
Ylinen
et
al.,
1989).
The
male
rats
(
N=
20)
were
castrated
at
the
age
of
28
days
and
after
5
weeks
were
used
in
the
tests.
Half
of
the
operated
and
10
intact
males
were
administered
estradiol
valerate
subcutaneously
500
ug/
kg
every
second
day
during
14
days
before
the
test.
Urine
was
collected
in
metabolism
cages
during
96
hr
after
a
single
intraperitoneal
PFOA
dose
(
50
mg/
kg).
Blood
samples
were
collected
by
cardiac
puncture.
Castration
and
administration
of
estradiol
to
the
male
rats
had
a
significant
stimulatory
effect
on
the
urinary
excretion
of
PFOA.
During
the
first
24
hours,
female
rats
excreted
72+­
5%
(
N=
6)
of
the
dose,
whereas
the
intact
males
excreted
only
9+­
4%
(
N=
6).
After
the
estradiol
treatment,
both
the
intact
and
castrated
males
excreted
PFOA
in
amounts
similar
to
females
(
61+­
19%
and
68+­
14%,
respectively).
The
castrated
males
without
estradiol
treatment
excreted
PFOA
in
urine
faster
than
the
intact
males
(
50+­
13%),
but
less
than
the
females
and
the
estrogen
treated
males.
At
the
end
of
the
test
(
96
hr),
the
concentration
of
PFOA
in
the
serum
of
intact
males
was
considerably
higher
(
17­
40
times)
than
in
the
serum
of
other
groups.
There
was
no
statistically
significant
difference
in
the
serum
concentrations
between
the
other
groups.
PFOA
was
similarly
bound
by
the
proteins
in
the
serum
of
males
and
females.

Vanden
Heuvel
et
al.
(
1992a)
investigated
whether
androgens
or
estrogens
are
involved
in
the
marked
sex­
differences
in
the
urinary
excretion
of
PFOA.
Castration
of
males
greatly
increased
(>
1­
fold)
the
elimination
of
14C­
PFOA
(
9.4
umol/
kg,
i.
p.)
in
urine,
demonstrating
that
a
factor
produced
by
the
testis
is
responsible
for
the
slow
elimination
of
PFOA
in
male
rats.
Castration
plus
17B­
estradiol
had
no
further
effect
on
PFOA
elimination
whereas
castration
plus
testosterone
replacement
at
the
physiological
level
reduced
PFOA
elimination
to
the
same
level
as
rats
with
intact
testis.
Thus,
in
male
rats,
testosterone
exerts
an
inhibitory
effect
on
renal
32
excretion
of
PFOA.
In
female
rats,
neither
ovariectomy
or
ovariectomy
plus
testosterone
affected
the
urinary
excretion
of
PFOA,
demonstrating
that
the
inhibitory
effect
of
testosterone
on
PFOA
renal
excretion
is
a
male­
specific
response.
Probenecid,
which
inhibits
the
renal
transport
system,
decreased
the
high
rate
of
PFOA
renal
excretion
in
castrated
males
but
had
no
effect
on
male
rats
with
intact
testis.

Hormonal
changes
during
pregnancy
do
not
appear
to
cause
a
change
in
the
rate
of
elimination
of
14C
after
oral
administration
of
a
single
dose
of
ammonium
14C­
PFOA
(
Gibson
and
Johnson,
1983).
At
8
or
9
days
after
conception,
four
pregnant
rats
and
2
nonpregnant
female
rats
were
dosed
(
mean
dose,
15
mg/
kg)
and
individual
urine
samples
were
collected
at
12,
24,
36,
and
48
hours
post
dose
and
analyzed
for
14C
content.
Essentially
all
of
the
14C
was
eliminated
via
urine
within
24
hours
for
both
groups
of
rats.

Feeding
of
cholestyramine
to
rats
enhanced
the
fecal
elimination
of
APFO
(
Johnson
et
al.
(
1984).
Male
rats
were
administered
APFO
(
2.1
mg/
ml)
dissolved
in
0.9%
NaCl
as
a
single
intravenous
dose
(
2
ml/
rat,
average
APFO
dose
13
mg/
kg).
At
14
days
post
dose,
the
mean
percentage
of
APFO
dose
eliminated
in
the
feces
of
cholestyramine­
treated
rats
(
43.2+­
5.5)
was
9.8­
fold
the
mean
percentage
of
dose
eliminated
in
feces
by
untreated
rats
(
4.4+­
1.0).
Excretion
in
urine
was
41%
for
treated
rats
and
67%
for
untreated
rats.

Kudo
et
al.
(
2002)
demonstrated
in
male
and
female
rats
that
renal
clearance
(
CLR)
of
PFOA
and
the
renal
mRNA
levels
of
specific
organic
anion
transporters
are
markedly
affected
by
sex
hormones.
The
biological
half­
life
of
PFOA
in
male
rats
was
found
to
be
70
times
longer
than
in
female
rats
and
this
difference
is
due
primarily
to
low
CLR
in
male
rats.
Castration
of
male
rats
caused
a
14­
fold
increase
in
CLR
of
PFOA.
The
elevated
PFOA
CLR
in
castrated
males
was
reduced
by
treating
them
with
testosterone.
Treatment
of
male
rats
with
estradiol
increased
the
CLR
of
PFOA.
In
female
rats,
ovariectomy
caused
a
significant
increase
in
CLR
of
PFOA,
which
was
reduced
by
estradiol
treatment.
Treatments
of
female
rats
with
testosterone
reduced
the
CLR
of
PFOA.
Treatment
with
probenecid,
a
known
inhibitor
of
organic
anion
transporters,
markedly
reduced
the
CLR
of
PFOA
in
male
rats,
castrated
male
rats,
and
female
rats.
To
identify
the
transporter
molecules
that
are
responsible
for
PFOA
transport
in
the
rat
kidney,
renal
mRNA
levels
of
specific
organic
anion
transporters
were
determined
in
male
and
female
rats
under
various
hormonal
states
and
compared
with
the
CLR
of
PFOA.
The
level
of
OAT2
mRNA
in
male
rats
was
only
13%
that
in
female
rats.
Castration
or
estradiol
treatment
increased
the
level
of
OAT2
mRNA
whereas
treatment
of
castrated
male
rats
with
testosterone
reduced
it.
Ovariectomy
of
female
rats
significantly
increased
the
level
of
OAT3
mRNA.
Multiple
regression
analysis
of
the
data
suggested
that
organic
anion
transporter
2
(
OAT2)
and
OAT3
are
responsible
for
urinary
elimination
of
PFOA
in
the
rat.

3.2
Epidemiology
Studies
3.2.1
Medical
Surveillance
Studies
from
the
Antwerp
and
Decatur
Plants
A
cross­
sectional
analysis
of
the
data
from
the
2000
medical
surveillance
program
at
the
Decatur
33
and
Antwerp
plants
was
undertaken
to
determine
if
there
were
any
associations
between
PFOA
and
hematology,
clinical
chemistries,
and
hormonal
parameters
of
volunteer
employees
(
Olsen,
et
al.,
2001e).
The
data
were
analyzed
for
all
employees
from
both
plant
locations.
Mean
PFOA
serum
levels
were
1.03
ppm
for
all
male
employees
at
the
Antwerp
plant
and
1.90
ppm
for
all
male
employees
at
the
Decatur
plant.
Male
production
employees
at
the
Decatur
plant
had
significantly
higher
(
p
<
.05)
mean
serum
levels
(
2.34
ppm)
than
those
at
the
Antwerp
plant
(
1.28
ppm).
Non­
production
employees
at
both
plants
had
mean
levels
below
1
ppm.
PFOA
serum
levels
were
higher
than
the
PFOS
serum
values
at
both
plants,
especially
the
Decatur
plant
where
serum
levels
are
higher
overall.
In
addition,
values
for
total
organic
fluorine
were
even
higher
than
the
PFOA
levels.

Multivariable
regression
analyses
were
conducted
to
adjust
for
possible
confounders
that
may
affect
the
results
of
the
clinical
chemistry
tests.
The
following
variables
were
included:
production
job
(
yes
or
no),
plant,
age,
body
mass
index
(
BMI),
cigarettes/
day,
drinks/
day
and
years
worked
at
the
plant.
A
positive
significant
association
was
reported
between
PFOA
and
cholesterol
(
p
=
.05)
and
PFOA
and
triglycerides
(
p
=
.002).
Age
was
also
significant
in
both
analyses.
Alcohol
consumed
per
day
was
significant
in
the
cholesterol
model,
while
BMI
and
cigarettes
smoked
per
day
was
significant
for
triglycerides.
When
both
PFOA
and
PFOS
were
included
in
the
analyses,
neither
reached
statistical
significance
in
the
cholesterol
model,
while
PFOA
remained
significant
(
p
=
.02)
in
the
triglycerides
model.
HDL
was
negatively
associated
with
PFOA
(
p
=
.04)
and
remained
significant
(
p
=
.04)
when
both
PFOA
and
PFOS
were
included
in
the
model.
A
positive
association
(
p
=
.01)
between
T3
and
PFOA
was
also
observed
and
remained
statistically
significant
(
p
=
.05)
when
PFOS
was
included
in
the
model.
BMI,
cigarettes/
day,
alcohol/
day
were
also
significant
in
the
model.
None
of
the
other
clinical
chemistry,
thyroid
or
hematology
measures
were
significantly
associated
with
PFOA
in
the
regression
model.

A
longitudinal
analysis
of
the
above
data
and
previous
medical
surveillance
results
was
performed
to
determine
whether
occupational
exposure
to
fluorochemicals
over
time
is
related
to
changes
in
clinical
chemistry
and
lipid
results
in
employees
of
the
Antwerp
and
Decatur
facilities
(
Olsen,
et
al.,
2001f).
The
clinical
chemistries
included:
cholesterol,
HDL,
triglycerides,
alkaline
phosphatase,
gamma
glutamyl
transferase
(
GGT),
aspartate
aminotransferase
(
AST),
alanine
aminotransferase
(
ALT),
total
and
direct
bilirubin.
Medical
surveillance
data
from
1995,
1997,
and
2000
were
analyzed
using
multivariable
regression.
The
plants
were
analyzed
using
3
subcohorts
that
included
those
who
participated
in
2
or
more
medical
exams
between
1995
and
2000.
A
total
of
175
male
employees
voluntarily
participated
in
the
2000
surveillance
and
at
least
one
other.
Only
41
employees
were
participants
in
all
3
surveillance
periods.

When
mean
serum
PFOA
levels
were
compared
by
surveillance
year,
PFOA
levels
in
the
employees
participating
in
medical
surveillance
at
the
Antwerp
plant
increased
between
1994/
95
and
1997
and
then
decreased
slightly
between
1997
and
2000.
At
the
Decatur
plant,
PFOA
serum
levels
decreased
between
1994/
95
and
1997
and
then
increased
between
1997
and
2000.
When
analyzed
using
mixed
model
multivariable
regression
and
combining
Antwerp
and
Decatur
employees,
there
was
a
statistically
significant
positive
association
between
PFOA
and
34
serum
cholesterol
(
p
=
.0008)
and
triglycerides
(
p
=
.0002)
over
time.
When
analyzed
by
plant
and
also
by
subcohort,
these
associations
were
limited
to
the
Antwerp
employees
(
p
=
.005)
and,
in
particular,
the
21
Antwerp
employees
who
participated
in
all
3
surveillance
years
(
p
=
.001).
However,
the
association
between
PFOA
and
triglycerides
was
also
statistically
significant
(
p
=
.02)
for
the
subgroup
in
which
employees
participated
in
biomonitoring
in
1994/
95
and
2000.
There
was
not
a
significant
association
between
PFOA
and
triglycerides
among
Decatur
workers.
There
were
no
significant
associations
between
PFOA
and
changes
over
time
in
HDL,
alkaline
phosphatase,
GGT,
AST,
ALT,
total
bilirubin,
and
direct
bilirubin.

There
are
several
limitations
to
the
2000
cross­
sectional
and
longitudinal
studies
including:
1)
serum
PFOA
levels
were
significantly
higher
at
the
Decatur
plant
than
at
the
Antwerp
plant,
2)
all
participants
were
volunteers,
3)
there
were
several
consistent
differences
in
clinical
chemistry
profiles
and
demographics
between
employees
of
the
Decatur
and
Antwerp
plants
(
Antwerp
employees
as
compared
to
Decatur
employees
had
lower
PFOA
serum
levels,
were
younger,
had
lower
BMIs,
worked
fewer
years,
had
higher
alcohol
consumption,
higher
mean
HDL
and
bilirubin
values,
lower
mean
triglyceride,
alkaline
phosphatase,
GGT,
AST,
and
ALT
values,
and
mean
thyroid
hormone
values
tended
to
be
higher),
4)
PFOS
and
other
perfluorinated
chemicals
are
also
present
in
these
plants,
5)
in
the
cross­
sectional
study,
plant
populations
cannot
be
compared
because
they
were
placed
into
quartiles
based
on
PFOS
serum
distributions
only
which
were
different
for
each
subgroup
and
not
applicable
to
PFOA,
6)
only
one
measurement
at
a
certain
point
in
time
was
collected
for
each
clinical
chemistry
test,
and
7)
PFOA
serum
levels
overall
have
been
increasing
over
time
in
these
employees.
In
addition,
in
the
longitudinal
study
only
a
small
number
of
employees
participated
in
all
3
sampling
periods
(
24%),
different
labs
and
analytical
techniques
for
PFOA
were
used
each
year,
and
female
employees
could
not
be
analyzed
because
of
the
small
number
of
participants.

3.2.2
Medical
Surveillance
Studies
from
the
Cottage
Grove
Plant
A
voluntary
medical
surveillance
program
was
offered
to
employees
of
the
Cottage
Grove,
Minnesota
plant
in
1993,
1995,
and
1997
(
n
=
111,
80
and
74
employees,
respectively)
(
Olsen,
et
al.,
1998b,
Olsen
et
al.,
2000).
The
clinical
chemistry
parameters
(
cholesterol,
hepatic
enzymes,
and
lipoprotein
levels)
used
in
the
longitudinal
and
cross­
sectional
studies
of
the
Antwerp
and
Decatur
plants
were
also
used
in
this
study.
In
addition,
in
1997
only,
cholecystokinin­
33
(
CCK)
was
also
measured
at
the
Cottage
Grove
plant.
CCK
levels
were
observed
because
certain
research
has
suggested
that
pancreas
acinar
cell
adenomas
seen
in
rats
exposed
to
PFOA
may
be
the
result
of
increased
CCK
levels
(
Obourn,
et
al.,
1997).

Only
male
employees
involved
in
PFOA
production
were
included
in
the
study.
Sixty­
eight
employees
were
common
to
the
1993
and
1995
sampling
periods,
21
were
common
between
1995
and
1997,
and
17
participated
in
all
three
surveillance
years.
Mean
serum
PFOA
levels
and
ranges
are
provided
in
Table
2
of
the
Biomonitoring
Section
of
this
report.
It
should
be
noted
that
Cottage
Grove
has
the
highest
serum
PFOA
levels
of
the
3
plants
studied.

Employees'
serum
PFOA
levels
were
stratified
into
3
categories
(<
1,
1­
<
10,
and
 
10
ppm),
35
chosen
to
provide
a
greater
number
of
employees
in
the
 
10
ppm
category.
As
employees'
mean
serum
PFOA
levels
increased,
no
statistically
significant
abnormal
liver
function
tests,
hypolipidemia,
or
cholestasis
were
observed
in
any
of
the
sampling
years.
Multivariable
regression
analyses
controlling
for
potential
confounders
(
age,
alcohol
consumption,
BMI,
and
cigarettes
smoked)
yielded
similar
results.
The
authors
also
reported
that
renal
function,
blood
glucose,
and
hematology
measures
were
not
associated
with
serum
PFOA
levels;
however,
these
data
were
not
provided
in
the
paper.

The
mean
CCK
value
reported
for
the
1997
sample
was
28.5
pg/
ml
(
range
8.8
­
86.7
pg/
ml).
The
means
in
the
2
serum
categories
<
10
ppm
were
at
least
50%
higher
than
in
the
 
10
ppm
category.
A
statistically
significant
(
p
=
.03)
negative
association
between
mean
CCK
levels
and
the
3
PFOA
serum
categories
was
observed.
A
scatter
plot
of
the
natural
log
of
CCK
and
PFOA
shows
that
all
but
2
CCK
values
are
within
the
assay's
reference
range
of
0
­
80
pg/
ml.
Both
of
these
employees
(
CCK
values
of
80.5
and
86.7
pg/
ml)
had
serum
PFOA
levels
less
than
10
ppm
(
0.6
and
5.6
ppm,
respectively).
A
multiple
regression
model
of
the
natural
log
of
CCK
and
serum
PFOA
levels
continued
to
display
a
negative
association
after
adjusting
for
potential
confounders.

The
cross­
sectional
design
is
a
limitation
of
this
study.
Only
17
subjects
were
common
to
all
3
sampling
years.
In
addition,
the
medical
surveillance
program
is
a
voluntary
one.
The
participation
rate
of
eligible
production
employees
decreased
from
approximately
70%
in
1993
to
50%
in
1997.
Also,
the
laboratory
reference
range
changed
substantially
for
ALT
in
1997.
Finally,
different
analytical
methods
were
used
to
measure
serum
PFOA.
Serum
PFOA
was
determined
by
electrospray
high­
performance
liquid
chromatography/
mass
spectrometry
in
1997,
but
by
thermospray
in
1993
and
1995.

An
earlier
medical
surveillance
study
on
workers
who
were
employed
in
the
1980'
s
was
conducted
at
the
Cottage
Grove
plant;
however,
total
serum
fluorine
was
measured
instead
of
PFOA
(
Gilliland
and
Mandel,
1996).
Based
on
animal
studies
that
reported
that
animals
exposed
to
PFOA
develop
hepatomegaly
and
alterations
in
lipid
metabolism,
a
cross­
sectional,
occupational
study
was
performed
to
determine
if
similar
effects
are
present
in
workers
exposed
to
PFOA.
In
a
PFOA
production
facility,
115
workers
were
studied
to
determine
whether
serum
PFOA
affected
their
cholesterol,
lipoproteins,
and
hepatic
enzymes.
Forty­
eight
workers
who
were
exposed
to
PFOA
from
1985­
1989
were
included
in
the
study
(
96%
participation
rate).
Sixty­
five
employees
who
either
volunteered
or
were
asked
to
participate,
were
included
in
the
unexposed
group.
These
employees
were
assumed
to
have
little
or
no
PFOA
exposure
based
on
their
job
description.
However,
when
serum
levels
were
analyzed,
it
was
noted
that
this
group
of
workers
had
PFOA
levels
much
greater
than
the
general
population.
Therefore,
instead
of
job
categories,
total
serum
fluorine
was
used
to
classify
workers
into
exposure
groups.

Total
serum
fluorine
was
used
as
a
surrogate
measure
for
PFOA.
Serum
PFOA
was
not
measured,
due
to
the
cost
of
analyzing
the
samples.
Blood
samples
were
analyzed
for
total
serum
fluorine,
serum
glutamyl
oxaloacetic
transaminase
(
SGOT
or
AST),
serum
glutamyl
pyruvic
transaminase
(
SGPT
or
ALT),
gamma
glutamyl
transferase
(
GGT),
cholesterol,
low­
36
density
lipoproteins
(
LDL),
and
high­
density
lipoproteins
(
HDL).
All
of
the
participants
were
placed
into
five
categories
of
total
serum
fluorine
levels:
<
1
ppm,
1­
3
ppm,
>
3
­
10
ppm,
>
10
­
15
ppm,
and
>
15
ppm.
The
range
of
the
serum
fluorine
values
was
0
to
26
ppm
(
mean
3.3
ppm).
Approximately
half
of
the
workers
fell
into
the
>
1
­
3
ppm
category,
while
23
had
serum
levels
<
1
ppm
and
11
had
levels
>
10
ppm.

There
were
no
significant
differences
between
exposure
categories
when
analyzed
using
univariate
analyses
for
cholesterol,
LDL,
and
HDL.
In
the
multivariate
analysis,
there
was
not
a
significant
association
between
total
serum
fluorine
and
cholesterol
or
LDL
after
adjusting
for
alcohol
consumption,
age,
BMI,
and
cigarette
smoking.
There
were
no
statistically
significant
differences
among
the
exposure
categories
of
total
serum
fluorine
for
AST,
ALT
and
GGT.
However,
increases
in
AST
and
ALT
occurred
with
increasing
total
serum
fluorine
levels
in
obese
workers
(
BMI
=
35
kg/
m2).
This
result
was
not
observed
when
PFOA
was
measured
directly
in
serum
of
workers
in
1993,
1995,
or
1997
surveillance
data
of
employees
of
the
Cottage
Grove
plant
(
Olsen,
et
al.,
2000).

Since
PFOA
was
not
measured
directly
and
there
is
no
exposure
information
provided
on
the
employees
(
eg.
length
of
employment/
exposure),
the
results
of
the
study
provide
limited
information.
The
authors
state
that
no
adverse
clinical
outcomes
related
to
PFOA
exposure
have
been
observed
in
these
employees;
however,
it
is
not
clear
that
there
has
been
follow­
up
of
former
employees.
In
addition,
the
range
of
results
reported
for
the
liver
enzymes
were
fairly
wide
for
many
of
the
exposure
categories,
indicating
variability
in
the
results.
Given
that
only
one
sample
was
taken
from
each
employee,
this
is
not
surprising.
It
would
be
much
more
helpful
to
have
several
samples
taken
over
time
to
ensure
their
reliability.
It
also
would
have
been
interesting
to
compare
the
results
of
the
workers
who
were
known
to
be
exposed
to
PFOA
to
those
who
were
originally
thought
not
to
be
exposed
to
see
if
there
were
any
differences
among
the
employees
in
these
groups.
There
were
more
of
the
"
unexposed"
employees
(
n
=
65)
participating
in
the
study
than
those
who
worked
in
PFOA
production
(
n
=
48).

3.2.3
Mortality
Studies
A
retrospective
cohort
mortality
study
was
performed
on
employees
at
the
Cottage
Grove,
MN
plant
which
produces
APFO
(
Gilliland
and
Mandel,
1993).
At
this
plant,
APFO
production
was
limited
to
the
Chemical
Division.
The
cohort
consisted
of
workers
who
had
been
employed
at
the
plant
for
at
least
6
months
between
January
1947
and
December
1983.
Death
certificates
of
all
of
the
workers
were
obtained
to
determine
cause
of
death.
There
was
almost
complete
follow­
up
(
99.5%)
of
all
of
the
study
participants.
The
exposure
status
of
the
workers
was
categorized
based
on
their
job
histories.
If
they
had
been
employed
for
at
least
1
month
in
the
Chemical
Division,
they
were
considered
exposed.
All
others
were
considered
to
be
not
exposed
to
PFOA.
The
number
of
months
employed
in
the
Chemical
Division
provided
the
cumulative
exposure
measurements.
Of
the
3537
(
2788
men
and
749
women)
employees
who
participated
in
this
study,
398
(
348
men
and
50
women)
were
deceased.
Eleven
of
the
50
women
and
148
of
the
348
men
worked
in
the
Chemical
Division,
and
therefore,
were
considered
exposed
to
PFOA.
37
Standardized
Mortality
Ratios
(
SMRs),
adjusted
for
age,
sex,
and
race
were
calculated
and
compared
to
U.
S.
and
Minnesota
white
death
rates
for
men.
For
women,
only
state
rates
were
available.
The
SMRs
for
males
were
stratified
for
3
latency
periods
(
10,
15,
and
20
years)
and
3
periods
of
duration
of
employment
(
5,
10,
and
20
years).

For
all
female
employees,
the
SMRs
for
all
causes
and
for
all
cancers
were
less
than
1.
The
only
elevated
(
although
not
significant)
SMR
was
for
lymphopoietic
cancer,
and
was
based
on
only
3
deaths.
When
exposure
status
was
considered,
SMRs
for
all
causes
of
death
and
for
all
cancers
were
significantly
lower
than
expected,
based
on
the
U.
S.
rates,
for
both
the
Chemical
Division
workers
and
the
other
employees
of
the
plant.

In
all
male
workers
at
the
plant,
the
SMRs
were
close
to
1
for
most
of
the
causes
of
death
when
compared
to
both
the
U.
S.
and
the
Minnesota
death
rates.
When
latency
and
duration
of
employment
were
considered,
there
were
no
elevated
SMRs.
When
employee
deaths
in
the
Chemical
Division
were
compared
to
Minnesota
death
rates,
the
SMR
for
prostate
cancer
for
workers
in
the
Chemical
Division
was
2.03
(
95%
CI
.55
­
4.59).
This
was
based
on
4
deaths
(
1.97
expected).
There
was
also
a
statistically
significant
association
with
length
of
employment
in
the
Chemical
Division
and
prostate
cancer
mortality.
Based
on
the
results
of
proportional
hazard
models,
the
relative
risk
for
a
1­
year
increase
in
employment
in
the
Chemical
Division
was
1.13
(
95%
CI
1.01
to
1.27).
It
rose
to
3.3
(
95%
CI
1.02
­
10.6)
for
workers
employed
in
the
Chemical
Division
for
10
years
when
compared
to
the
other
employees
in
the
plant.
The
SMR
for
workers
not
employed
in
the
Chemical
Division
was
less
than
expected
for
prostate
cancer
(.
58).

An
update
of
this
study
was
conducted
to
include
the
death
experience
of
employees
through
1997
(
Alexander,
2001a).
The
cohort
consisted
of
3992
workers.
The
eligibility
requirement
was
increased
to
1
year
of
employment
at
the
Cottage
Grove
plant,
and
the
exposure
categories
were
changed
to
be
more
specific.
Workers
were
placed
into
3
exposure
groups
based
on
job
history
information:
definite
PFOA
exposure
(
n
=
492,
jobs
where
cell
generation,
drying,
shipping
and
packaging
of
PFOA
occurred
throughout
the
history
of
the
plant);
probable
PFOA
exposure
(
n
=
1685,
other
chemical
division
jobs
where
exposure
to
PFOA
was
possible
but
with
lower
or
transient
exposures);
and
not
exposed
to
fluorochemicals
(
n
=
1815,
primarily
nonchemical
division
jobs).

In
this
new
cohort,
607
deaths
were
identified:
46
of
these
deaths
were
in
the
PFOA
exposure
group,
267
in
the
probable
exposure
group,
and
294
in
the
non­
exposed
group.
When
all
employees
were
compared
to
the
state
mortality
rates,
SMRs
were
less
than
1
or
only
slightly
higher
for
all
of
the
causes
of
death
analyzed.
None
of
the
SMRs
were
statistically
significant
at
p
=
.05.
The
highest
SMR
reported
was
for
bladder
cancer
(
SMR
=
1.31,
95%
CI
=
0.42
 
3.05).
Five
deaths
were
observed
(
3.83
expected).

A
few
SMRs
were
elevated
for
employees
in
the
definite
PFOA
exposure
group:
2
deaths
from
cancer
of
the
large
intestine
(
SMR
=
1.67,
95%
CI
=
0.02
 
6.02),
1
from
pancreatic
cancer
38
(
SMR
=
1.34,
95%
CI
=
0.03
 
7.42),
and
1
from
prostate
cancer
(
SMR
=
1.30,
95%
CI
=
0.03
 
7.20).
In
addition,
employees
in
the
definite
PFOA
exposure
group
were
2.5
times
more
likely
to
die
from
cerebrovascular
disease
(
5
deaths
observed,
1.94
expected;
95%
CI
=
0.84
 
6.03).

In
the
probable
exposure
group,
3
SMRs
should
be
noted:
cancer
of
the
testis
and
other
male
genital
organs
(
SMR
=
2.75,
95%
CI
=
0.07
 
15.3);
pancreatic
cancer
(
SMR
=
1.24,
95%
CI
=
0.45
 
2.70);
and
malignant
melanoma
of
the
skin
(
SMR
=
1.42,
95%
CI
=
0.17
 
5.11).
Only
1,
6,
and
2
cases
were
observed,
respectively.
The
SMR
for
prostate
cancer
in
this
group
was
0.86
(
95%
CI
=
0.28
 
2.02)
(
n
=
5).

There
were
no
notable
excesses
in
SMRs
in
the
non­
exposed
group,
except
for
cancer
of
the
bladder
and
other
urinary
organs.
Four
cases
were
observed
and
only
1.89
were
expected
(
95%
CI
=
0.58
 
5.40).

It
is
difficult
to
interpret
the
results
of
the
prostate
cancer
deaths
between
the
first
study
and
the
update
because
the
exposure
categories
were
modified
in
the
update.
Only
1
death
was
reported
in
the
definite
exposure
group
and
5
were
observed
in
the
probable
exposure
group.
All
of
these
deaths
would
have
been
placed
in
the
chemical
plant
employees
exposure
group
in
the
first
study.
The
number
of
years
that
these
employees
worked
at
the
plant
and/
or
were
exposed
to
PFOA
was
not
reported.
This
is
important
because
even
1
prostate
cancer
death
in
the
definite
PFOA
exposure
group
resulted
in
an
elevated
SMR
for
the
group.
Therefore,
if
any
of
the
employees'
exposures
were
misclassified,
the
results
of
the
analysis
could
be
altered
significantly.

The
excess
mortality
in
cerebrovascular
disease
noted
in
employees
in
the
definite
exposure
group
was
further
analyzed
based
on
number
of
years
of
employment
at
the
plant.
Three
of
the
5
deaths
occurred
in
workers
who
were
employed
in
jobs
with
definite
PFOA
exposure
for
more
than
5
years
but
less
than
10
years
(
SMR
=
15.03,
95%
CI
=
3.02
 
43.91).
The
other
2
occurred
in
employees
with
less
than
1
year
of
definite
exposure.
The
SMR
was
6.9
(
95%
CI
=
1.39
 
20.24)
for
employees
with
greater
than
5
years
of
definite
PFOA
exposure.
In
order
to
confirm
that
the
results
regarding
cerebrovascular
disease
were
not
an
artifact
of
death
certificate
coding,
regional
mortality
rates
were
used
for
the
reference
population.
The
results
did
not
change.
When
these
deaths
were
further
analyzed
by
cumulative
exposure
(
time­
weighted
according
to
exposure
category),
workers
with
27
years
of
exposure
in
probable
PFOA
exposed
jobs
or
those
with
9
years
of
definite
PFOA
exposure
were
3.3
times
more
likely
to
die
of
cerebrovascular
disease
than
the
general
population.
A
dose­
response
relationship
was
not
observed
with
years
of
exposure.

It
is
difficult
to
compare
the
results
of
the
first
and
second
mortality
studies
at
the
Cottage
Grove
plant
since
the
exposure
categories
were
modified.
Although
the
potential
for
exposure
misclassification
was
certainly
more
likely
in
the
first
study,
it
may
still
have
occurred
in
the
update
as
well.
It
is
difficult
to
judge
the
reliability
of
the
exposure
categories
that
were
defined
without
measured
exposures.
Although
serum
PFOA
measurements
were
considered
in
the
exposure
matrix
developed
for
the
update,
they
were
not
directly
used.
In
the
second
study,
the
39
chemical
plant
employees
were
sub­
divided
into
PFOA­
exposed
groups,
and
the
film
plant
employees
essentially
remained
in
the
"
non­
exposed"
group.
This
was
an
effort
to
more
accurately
classify
exposures;
however,
these
new
categories
do
not
take
into
account
duration
of
exposure
or
length
of
employment.
Another
limitation
to
this
study
is
that
17
death
certificates
were
not
located
for
deceased
employees
and
therefore
were
not
included
in
the
study.
The
inclusion
or
exclusion
of
these
deaths
could
change
the
analyses
for
the
causes
of
death
that
had
a
small
number
of
cases.
Follow
up
of
worker
mortality
at
Cottage
Grove
(
and
Decatur)
needs
to
continue.
Although
there
were
more
than
200
additional
deaths
included
in
this
analysis,
it
is
a
small
number
and
the
cohort
is
still
relatively
young.
Given
the
results
of
studies
on
fluorochemicals
in
both
animals
and
humans,
further
analysis
is
warranted.

3.2.4
Hormone
Study
Endocrine
effects
have
been
associated
with
PFOA
exposure
in
animals;
therefore,
medical
surveillance
data,
including
hormone
testing,
from
employees
of
the
Cottage
Grove,
Minnesota
plant
were
analyzed
(
Olsen,
et
al.,
1998a).
PFOA
serum
levels
were
obtained
for
volunteer
workers
in
1993
(
n
=
111)
and
1995
(
n
=
80).
Sixty­
eight
employees
were
common
to
both
sampling
periods.
In
1993,
the
range
of
PFOA
was
0­
80
ppm
(
although
80
ppm
was
the
limit
of
detection
that
year,
so
it
could
have
been
higher)
and
0­
115
ppm
in
1995
using
thermospray
mass
spectrophotometry
assay.
Eleven
hormones
were
assayed
from
the
serum
samples.
They
were:
cortisol,
dehydroepiandrosterone
sulfate
(
DHEAS),
estradiol,
FSH,
17
gammahydroxyprogesterone
(
17­
HP),
free
testosterone,
total
testosterone,
LH,
prolactin,
thyroidstimulating
hormone
(
TSH)
and
sex
hormone­
binding
globulin
(
SHBG).
Employees
were
placed
into
4
exposure
categories
based
on
their
serum
PFOA
levels:
0­
1
ppm,
1­
<
10
ppm,
10­
<
30
ppm,
and
>
30
ppm.
Statistical
methods
used
to
compare
PFOA
levels
and
hormone
values
included:
multivariable
regression
analysis,
ANOVA,
and
Pearson
correlation
coefficients.

PFOA
was
not
highly
correlated
with
any
of
the
hormones
or
with
the
following
covariates:
age,
alcohol
consumption,
BMI,
or
cigarettes.
Most
of
the
employees
had
PFOA
serum
levels
less
than
10
ppm.
In
1993,
only
12
employees
had
serum
levels
>
10
ppm,
and
15
in
1995.
However,
these
levels
ranged
from
approximately
10
ppm
to
over
114
ppm.
There
were
only
4
employees
in
the
>
30
ppm
PFOA
group
in
1993
and
only
5
in
1995.
Therefore,
it
is
likely
that
there
was
not
enough
power
to
detect
differences
in
either
of
the
highest
categories.
The
mean
age
of
the
employees
in
the
highest
exposure
category
was
the
lowest
in
both
1993
and
1995
(
33.3
years
and
38.2
years,
respectively).
Although
not
significantly
different
from
the
other
categories,
BMI
was
slightly
higher
in
the
highest
PFOA
category.

Estradiol
was
highly
correlated
with
BMI
(
r
=
.41,
p
<
.001
in
1993,
and
r
=
.30,
p
<
.01
in
1995).
In
1995,
all
5
employees
with
PFOA
levels
>
30
ppm
had
BMIs
>
28,
although
this
effect
was
not
observed
in
1993.
Estradiol
levels
in
the
>
30
ppm
group
in
both
years
were
10%
higher
than
the
other
PFOA
groups;
however,
the
difference
was
not
statistically
significant.
The
authors
postulate
that
the
study
may
not
have
been
sensitive
enough
to
detect
an
association
between
PFOA
and
estradiol
because
measured
serum
PFOA
levels
were
likely
below
the
observable
effect
levels
suggested
in
animal
studies
(
55
ppm
PFOA
in
the
CD
rat).
Only
3
40
employees
in
this
study
had
PFOA
serum
levels
this
high.
They
also
suggest
that
the
higher
estradiol
levels
in
the
highest
exposure
category
could
suggest
a
threshold
relationship
between
PFOA
and
estradiol.

Free
testosterone
was
highly
correlated
with
age
in
both
1993
and
1995.
The
authors
did
not
report
a
negative
association
between
PFOA
serum
levels
and
testosterone.
There
were
no
statistically
significant
trends
noted
for
PFOA
and
either
bound
or
free
testosterone.
However,
17­
HP,
a
precursor
of
testosterone,
was
highest
in
the
>
30
ppm
PFOA
group
in
both
1993
and
1995.
In
1995,
PFOA
was
significantly
associated
with
17­
HP
in
regression
models
adjusted
for
possible
confounders.
However,
the
authors
state
that
this
association
was
based
on
the
results
of
one
employee
(
data
were
not
provided
in
the
report).
There
were
no
significant
associations
between
PFOA
and
cortisol,
DHEAS,
FSH,
LH,
and
SHBG.

There
are
several
design
issues
that
should
be
noted
when
evaluating
the
results
of
this
study.
First,
although
there
were
2
study
years
(
1993
and
1995),
the
populations
were
not
independent.
Sixty­
eight
employees
participated
in
both
years.
Second,
there
were
31
fewer
employees
who
participated
in
the
study
in
1995,
thus
reducing
the
power
of
the
study.
There
were
also
very
few
employees
in
either
year
with
serum
PFOA
levels
greater
than
10
ppm.
Third,
the
crosssectional
design
of
the
study
does
not
allow
for
analysis
of
temporality
of
an
association.
Since
the
half­
life
of
PFOA
is
at
least
1
year,
the
authors
suggest
that
it
is
possible
that
there
may
be
some
biological
accommodation
to
the
effects
of
PFOA.
Fourth,
only
one
sample
was
taken
for
each
hormone
for
each
of
the
study
years.
In
order
to
get
more
accurate
measurements
for
some
of
the
hormones,
pooled
blood
taken
in
a
short
time
period
should
have
been
used
for
each
participant.
Fifth,
some
of
the
associations
that
were
measured
in
this
study
were
done
based
on
the
results
of
an
earlier
paper
that
linked
PFOA
with
increased
estradiol
and
decreased
testosterone
levels.
However,
total
serum
organic
fluorine
was
measured
in
that
study
instead
of
PFOA,
making
it
difficult
to
compare
the
results.
Finally,
there
may
have
been
some
measurement
error
of
some
of
the
confounding
variables.

3.2.5
Study
on
Episodes
of
Care
(
Morbidity)

In
order
to
gain
additional
insight
into
the
effects
of
fluorochemical
exposure
on
workers'
health,
an
"
episode
of
care"
analysis
was
undertaken
at
the
Decatur
plant
to
screen
for
morbidity
outcomes
that
may
be
associated
with
long­
term,
high
exposure
to
fluorochemicals
(
Olsen
et
al.,
2001g).
An
"
episode
of
care"
is
a
series
of
health
care
services
provided
from
the
start
of
a
particular
disease
or
condition
until
solution
or
resolution
of
that
problem.
Episodes
of
care
were
identified
in
employees'
health
claims
records
using
Clinical
Care
Groups
(
CCG)
software.
All
inpatient
and
outpatient
visits
to
health
care
providers,
procedures,
ancillary
services
and
prescription
drugs
used
in
the
diagnosis,
treatment,
and
management
of
over
400
diseases
or
conditions
were
tracked.

Episodes
of
care
were
analyzed
for
652
chemical
employees
and
659
film
plant
employees
who
worked
at
the
Decatur
plant
for
at
least
1
year
between
January
1,
1993
and
December
31,
1998.
Based
on
work
history
records,
employees
were
placed
into
different
comparison
groups:
41
Group
A
consisted
of
all
film
and
chemical
plant
workers;
Group
B
had
employees
who
only
worked
in
either
the
film
or
chemical
plant;
Group
C
consisted
of
employees
who
worked
in
jobs
with
high
POSF
exposures;
and
Group
D
had
employees
who
worked
in
high
exposures
in
the
chemical
plant
for
10
years
or
more
prior
to
the
onset
of
the
study.
Film
plant
employees
were
considered
to
have
little
or
no
fluorochemical
exposure,
while
chemical
plant
employees
were
assumed
to
have
the
highest
exposures.

Ratios
of
observed
to
expected
episodes
of
care
were
calculated
for
each
plant.
Expected
numbers
were
based
on
3M's
employee
population
experience
using
indirect
standardization
techniques.
A
ratio
of
the
chemical
plant's
observed
to
expected
experience
divided
by
the
film
plant's
observed
to
expected
experience
was
calculated
to
provide
a
relative
risk
ratio
for
each
episode
of
care
(
RREpC).
95%
confidence
intervals
were
calculated
for
each
RREpC.
Episodes
of
care
that
were
of
greatest
interest
were
those
which
had
been
reported
in
animal
or
epidemiologic
literature
on
PFOS
and
PFOA:
liver
and
bladder
cancer,
endocrine
disorders
involving
the
thyroid
gland
and
lipid
metabolism,
disorders
of
the
liver
and
biliary
tract,
and
reproductive
disorders.

The
only
increased
risk
of
episodes
for
these
conditions
of
a
priori
interest
were
for
neoplasms
of
the
male
reproductive
system
and
for
the
overall
category
of
cancers
and
benign
growths
(
which
included
cancer
of
the
male
reproductive
system).
There
was
an
increased
risk
of
episodes
for
the
overall
cancer
category
for
all
4
comparison
groups.
The
risk
ratio
was
greatest
in
the
group
of
employees
with
the
highest
and
longest
exposures
to
fluorochemicals
(
RREpC
=
1.6,
95%
CI
=
1.2
 
2.1).
Increased
risk
of
episodes
in
long­
time,
high­
exposure
employees
also
was
reported
for
male
reproductive
cancers
(
RREpC
=
9.7,
95%
CI
=
1.1
­
458).
It
should
be
noted
that
the
confidence
interval
is
very
wide
for
male
reproductive
cancers
and
the
sub­
category
of
prostate
cancer.
Five
episodes
of
care
were
observed
for
reproductive
cancers
in
chemical
plant
employees
(
1.8
expected),
of
which
4
were
prostate
cancers.
One
episode
of
prostate
cancer
was
observed
in
film
plant
employees
(
3.4
expected).
This
finding
should
be
noted
because
an
excess
in
prostate
cancer
mortality
was
observed
in
the
Cottage
Grove
plant
mortality
study
when
there
were
only
2
exposure
categories
(
chemical
division
employees
and
non­
chemical
division
employees).
The
update
of
the
study
sub­
divided
the
chemical
plant
employees
and
did
not
confirm
this
finding
when
exposures
were
divided
into
definitely
exposed
and
probably
exposed
employees.

There
was
an
increased
risk
of
episodes
for
neoplasms
of
the
gastrointestinal
tract
in
the
high
exposure
group
(
RREpC
=
1.8,
95%
CI
=
1.2­
3.0)
and
the
long­
term
employment,
high
exposure
group
(
RREpC
=
2.9,
95%
CI
=
1.7
 
5.2).
Most
of
the
episodes
were
attributable
to
benign
colonic
polyps.
Similar
numbers
of
episodes
were
reported
in
film
and
chemical
plant
employees.

In
the
entire
cohort,
only
1
episode
of
care
was
reported
for
liver
cancer
(
0.6
expected)
and
1
for
bladder
cancer
(
1.5
expected).
Both
occurred
in
film
plant
employees.
Only
2
cases
of
cirrhosis
of
the
liver
were
observed
(
0.9
expected),
both
in
the
chemical
plant.
There
was
a
greater
risk
of
lower
urinary
tract
infections
in
chemical
plant
employees,
but
they
were
mostly
due
to
recurring
42
episodes
of
care
by
the
same
employees.
It
is
difficult
to
draw
any
conclusions
about
these
observations,
given
the
small
number
of
episodes
reported.

Chemical
plant
employees
in
the
high
exposure,
long­
term
employment
group
were
2
½
times
more
likely
to
seek
care
for
disorders
of
the
biliary
tract
than
their
counterparts
in
the
film
plant
(
RREpC
=
2.6,
95%
CI
=
1.2
­
5.5).
Eighteen
episodes
of
care
were
observed
in
chemical
plant
employees
and
14
in
film
plant
workers.
The
sub­
categories
that
influenced
this
observation
were
episodes
of
cholelithiasis
with
acute
cholecystitis
and
cholelithiasis
with
chronic
or
unspecified
cholecystitis.
Most
of
the
observed
cases
occurred
in
chemical
plant
employees.

Risk
ratios
of
episodes
of
care
for
endocrine
disorders,
which
included
sub­
categories
of
thyroid
disease,
diabetes,
hyperlipidemia,
and
other
endocrine
or
nutritional
disorders,
were
not
elevated
in
the
comparison
groups.
Conditions
which
were
not
identified
a
priori
but
which
excluded
the
null
hypothesis
in
the
95%
confidence
interval
for
the
high
exposure,
long­
term
employment
group
included:
disorders
of
the
pancreas,
cystitis,
and
lower
urinary
tract
infections.

The
results
of
this
study
only
should
be
used
for
hypothesis
generation.
Although
the
episode
of
care
design
allowed
for
a
direct
comparison
of
workers
with
similar
demographics
but
different
exposures,
there
are
many
limitations
to
this
design.
The
limitations
include:
1)
episodes
of
care
are
reported,
not
disease
incidence,
2)
the
data
are
difficult
to
interpret
because
a
large
RREpC
may
not
necessarily
indicate
high
risk
of
incidence
of
disease,
3)
many
of
the
risk
ratios
for
episodes
of
care
had
very
wide
confidence
intervals,
thereby
providing
unstable
results,
4)
the
analysis
was
limited
to
6
years,
5)
the
utilization
of
health
care
services
may
reflect
local
medical
practice
patterns,
6)
individuals
may
be
counted
more
than
once
in
the
database
because
they
can
be
categorized
under
larger
or
smaller
disease
classifications,
7)
episodes
of
care
may
include
the
same
individual
several
times,
8)
not
all
employees
were
included
in
the
database,
such
as
those
on
long­
term
disability,
9)
the
analysis
may
be
limited
by
the
software
used,
which
may
misclassify
episodes
of
care,
10)
the
software
may
assign
2
different
diagnoses
to
the
same
episode,
and
11)
certain
services,
such
as
lab
procedures
may
not
have
been
reported
in
the
database.

3.3
Acute
Toxicity
Studies
in
Animals
3.3.1
Oral
Studies
The
acute
oral
toxicity
of
APFO
was
tested
in
male
and
female
rats
in
three
studies.
Death
occurred
at
concentrations
 
464
mg/
kg
(
Internat'l
Res
and
Dev
Corp.,
1978).
Abnormal
findings
upon
necropsy
(
kidney,
stomach,
uterus)
were
observed
(
Glaza,
1997)
at
500
mg/
kg
(
higher
concentrations
were
not
tested).
Clinical
signs
of
toxicity
observed
in
these
three
studies
included
the
following:
red­
stained
face,
stained
urogenital
area,
wet
urogenital
area,
hypoactivity,
hunched
posture,
staggered
gait,
excessive
salivation,
ptosis,
piloerection,
decreased
limb
tone,
ataxia,
corneal
opacity,
and
hypothermic
to
touch.

In
one
study
(
Internat'l
Res
and
Dev
Corp.,
1978),
the
oral
LD50
values
for
Charles
River
CD
43
rats
were
680
mg/
kg
(
399
 
1157
mg/
kg
95%
confidence
limit)
for
males;
430
mg/
kg
(
295
 
626
mg/
kg
95%
confidence
limit)
for
females;
and
540
mg/
kg
(
389
 
749
mg/
kg
95%
confidence
limit)
for
males
and
females.
The
remaining
two
studies
provided
LD50
values
of
(
1)
>
500
mg/
kg
for
male
Crl:
CD(
SD)
BR
rats,
and
250­
500
mg/
kg
for
female
Crl:
CD(
SD)
BR
rats
(
Glaza,
1997);
and
(
2)
<
1000
mg/
kg
for
male
and
female
Sherman­
Wistar
rats
(
3M
Company,
1976b).

3.3.2
Inhalation
Studies
The
acute
inhalation
toxicity
of
APFO
was
tested
in
male
and
female
Sprague­
Dawley
rats,
at
a
dose
level
of
18.6
mg/
L
(
nominal
concentration),
and
exposure
duration
of
one
hour.
Signs
of
toxicity
during
and
up
to
14
days
after
the
exposure
period,
included
the
following:
excessive
salivation,
excessive
lacrimation,
decreased
activity,
labored
breathing,
gasping,
closed
eyes,
mucoid
nasal
discharge,
irregular
breathing,
red
nasal
discharge,
yellow
staining
of
the
anogenital
fur,
dry
and
moist
rales,
red
material
around
the
eyes,
and
body
tremors.
Upon
necropsy,
lung
discoloration
was
observed
in
a
higher
than
normal
incidence
of
rats
(
8/
10).
Based
on
the
study
results,
the
test
substance
was
not
fatal
to
rats
at
a
nominal
exposure
concentration
of
18.6
mg/
L
and
exposure
duration
of
one
hour
(
Bio/
dynamics,
Inc.
1979).

3.3.3
Dermal
Studies
The
acute
dermal
toxicity
of
APFO
was
tested
in
male
and
female
Hra(
NZW)
SPF
rabbits,
at
a
dose
level
of
2000
mg/
kg,
and
a
24­
hour
exposure
period.
All
animals
appeared
normal
and
exhibited
body
weight
gain
throughout
the
study,
with
the
exception
of
one
male
that
lost
weight
during
the
first
week.
Dermal
irritation
consisted
of
slight
to
moderate
erythema,
edema,
and
atonia;
slight
desquamation;
coriaceousness;
and
fissuring.
No
visible
lesions
were
observed
upon
necropsy.
The
dermal
LD50
in
rabbits
was
determined
to
be
greater
than
2000
mg/
kg
(
Glaza,
1995).

3.3.4
Eye
Irritation
Studies
The
eye
irritation
potential
of
APFO
was
tested
in
albino
rabbits,
at
a
dose
level
of
0.1
gram.
In
two
of
three
studies,
APFO
was
determined
to
be
a
primary
ocular
irritant.
In
the
studies
in
which
APFO
was
found
to
be
a
primary
ocular
irritant,
APFO
was
left
in
contact
with
the
eye
for
7
days,
then
rinsed,
or
not
rinsed.
Irritation
scores
varied
during
the
observation
period.
Irritation
scores
of
the
conjunctivae,
iris,
and
cornea
ranged
from
2
 
4
in
one
study
(
Biosearch,
Inc.
1976)
and
from
2
 
10
in
the
other
study
(
3M
Company,
1976a).
In
both
studies,
irritation
remained
evident
for
the
duration
of
the
observation
period
(
7­
days
post­
exposure).
In
the
study
in
which
APFO
was
determined
to
be
a
non­
irritant
(
Gabriel),
the
test
substance
was
left
in
contact
with
the
eye
for
5
or
30
seconds,
and
then
the
eyes
were
rinsed.
In
this
study,
positive
scores
were
reported
for
conjunctivae
irritation
for
up
to
7­
days
post­
exposure,
so
the
author's
negative
conclusion
for
ocular
irritancy
is
problematic.
44
3.3.5
Skin
Irritation
Studies
The
skin
irritation
potential
of
APFO
was
tested
in
albino
rabbits
in
two
studies,
at
a
dose
level
of
0.5
grams,
under
occluded
test
conditions.
In
one
study
(
Riker
Laboratories,
Inc.
1983),
APFO
produced
irreversible
tissue
damage
in
female
rabbits,
following
a
3­
minute,
1­
hour,
and
4­
hour
contact
period.
Moderate
erythema
and
edema,
as
well
as
chemical
burn,
eschar,
and
necrosis,
were
observed
following
all
three
contact
periods.
An
endpoint
was
not
achieved
in
this
study
due
to
extreme
irritation
following
each
contact
period.
In
contrast
in
the
second
study
(
Gabriel),
APFO
was
reported
as
a
non­
irritant
of
skin
after
an
exposure
period
of
24
or
72
hours,
based
on
primary
irritation
scores
of
zero.

3.4
Mutagenicity
Studies
APFO
was
tested
twice
(
Lawlor,
1995;
1996)
for
its
ability
to
induce
mutation
in
the
Salmonella
 
E.
coli/
mammalian­
microsome
reverse
mutation
assay.
The
tests
were
performed
both
with
and
without
metabolic
activation.
A
single
positive
response
seen
at
one
dose
level
in
S.
typhimurium
TA1537
when
tested
without
metabolic
activation
was
not
reproducible.
APFO
did
not
induce
mutation
in
either
S.
typhimurium
or
E.
coli
when
tested
either
with
or
without
mammalian
activation.
APFO
did
not
induce
chromosomal
aberrations
in
human
lymphocytes
when
tested
with
and
without
metabolic
activation
up
to
cytotoxic
concentrations
(
Murli,
1996c;
NOTOX,
2000).
Sadhu
(
2002)
recently
reported
that
APFO
did
not
induce
gene
mutation
when
tested
with
or
without
metabolic
activation
in
the
K­
1
line
of
Chinese
hamster
ovary
(
CHO)
cells
in
culture.

Murli
(
1996b)
tested
APFO
twice
for
its
ability
to
induce
chromosomal
aberrations
in
CHO
cells.
In
the
first
assay,
APFO
induced
both
chromosomal
aberrations
and
polyploidy
in
both
the
presence
and
absence
of
metabolic
activation.
In
the
second
assay,
no
significant
increases
in
chromosomal
aberrations
were
observed
without
activation.
However,
when
tested
with
metabolic
activation,
APFO
induced
significant
increases
in
chromosomal
aberrations
and
in
polyploidy
(
Murli,
1996b).

APFO
was
tested
in
a
cell
transformation
and
cytotoxicity
assay
conducted
in
C3H
10T
½
mouse
embryo
fibroblasts.
The
cell
transformation
was
determined
as
both
colony
transformation
and
foci
transformation
potential.
There
was
no
evidence
of
transformation
at
any
of
the
dose
levels
tested
in
either
the
colony
or
foci
assay
methods
(
Garry
&
Nelson,
1981).

APFO
was
tested
twice
in
the
mouse
micronucleus
assay.
APFO
did
not
induce
any
significant
increases
in
micronuclei
and
was
considered
negative
under
the
conditions
of
this
assay
(
Murli,
1996a).

3.5
Subchronic
Toxicity
Studies
in
Animals
Subchronic
toxicity
studies
have
been
conducted
in
rats,
mice,
rhesus
monkeys
and
cynomolgus
monkeys.
Two
unpublished
28­
day
feeding
studies
were
performed
at
Industrial
Bio­
Test
45
Laboratories,
Inc.
(
Metrick
and
Marias,
1977
and
Christopher
and
Marias,
1977).
In
both
rats
and
mice
the
liver
was
the
target
organ.
In
rats,
males
had
more
pronounced
hepatotoxicity
and
histopathologic
effects
than
females.
Three
90­
day
subchronic
toxicity
studies
have
been
conducted
in
rats
(
Goldenthal,
1978a,
Palazzolo,
1993)
and
rhesus
monkeys
(
Goldenthal,
1978b).
Thomford
(
2001a,
b)
recently
conducted
a
range­
finding
and
a
6­
month
toxicity
study
in
cynomolgus
monkeys.
In
addition,
chronic/
carcinogenicity
studies
have
been
conducted
which
are
described
in
section
3.8.
In
all
species,
the
liver
is
the
main
target
organ.
In
rats,
males
had
more
pronounced
hepatotoxicity
and
histopathologic
effects
than
females,
presumably
because
of
the
gender
difference
in
elimination
of
APFO.

In
a
28­
day
study
of
ChR­
CD
albino
rats,
eight
randomly
assigned
groups
of
five
males
and
five
females
were
studied
(
Metrick
and
Marias,
1977).
After
rats
were
allowed
to
acclimate
for
a
week
in
individual
cages
they
then
received
similar
feed
containing
0,
30,
100,
300,
1000,
3000,
10,000,
or
30,000
ppm
APFO
for
28
days.
The
animals
were
observed
daily
and
body
weights
and
food
consumption
were
recorded
weekly.
Animals
that
died
during
the
study
were
examined
for
gross
pathology,
as
were
surviving
animals
at
28
days.
It
is
stated
that
the
study
included
a
complete
examination
of
gross
pathology
and
a
complete
set
of
tissues
and
organs
were
examined,
but
the
specific
list
is
not
supplied.
Livers
were
weighed
to
determine
relative
organ
weight
then
stained
for
histopathologic
examination.

All
animals
in
the
10,000
and
30,000­
ppm
groups
died
before
the
end
of
the
first
week.
There
were
no
premature
deaths
or
other
clinical
signs
of
toxicity
in
the
other
groups.
Body
weight
gains
were
reduced
in
the
groups
receiving
1000
or
more
ppm.
Slight
reductions
in
body
weight
gain
were
also
observed
in
males
exposed
to
300
ppm
and
males
and
females
fed
100
ppm.
Reduced
food
intake
was
observed
in
rats
fed
1000
ppm
or
higher
in
a
dose­
related
manner.
Relative
liver
weights
were
increased
in
males
fed
30
ppm
or
more
and
females
fed
300
ppm
or
more.
Gross
pathological
exam
did
not
reveal
treatment­
related
effects
in
kidneys
or
other
organs
besides
livers.
Focal
to
multifocal
cytoplasmic
enlargement
of
hepatocytes
was
noted
in
animals
fed
300
ppm,
and
multifocal
to
diffuse
enlargement
of
hepatocytes
was
noted
in
animals
fed
1000
ppm
or
higher.
These
effects
were
more
pronounced
in
males
(
Metrick
and
Marias.
1977).

In
a
28­
day
study
of
Charles
River­
CD
albino
mice,
eight
randomly
assigned
groups
of
five
males
and
five
females
were
studied
(
Christopher
and
Marisa,
1977).
After
mice
were
allowed
to
acclimate
for
8
days
in
individual
cages
they
then
received
similar
feed
containing
0,
30,
100,
300,
1000,
3000,
10,000,
or
30,000
ppm
of
APFO
for
28
days.
The
animals
were
observed
daily
and
body
weights
and
food
consumption
were
recorded
weekly.
Animals
that
died
during
the
study
were
examined
for
gross
pathology,
as
were
surviving
animals
at
28
days.
It
is
stated
the
study
included
a
complete
examination
of
gross
pathology
and
a
representative
set
of
tissues
and
organs
were
examined,
but
the
specific
list
is
not
supplied.
Livers
were
weighed
to
determine
relative
organ
weight
then
stained
for
histopathologic
examination.

All
animals
in
the
1000­
ppm
and
higher
groups
died
before
the
end
of
day
9.
The
entire
300­
ppm
group
died
within
26
days
except
1
male.
One
animal
in
each
of
the
30
and
100­
ppm
groups
46
died
prematurely.
Clinical
signs
were
observed
in
mice
exposed
to
100
ppm
and
higher
doses
of
PFOA.
At
100
ppm
some
animals
exhibited
cyanosis
on
days
10
and
11
of
testing,
but
appeared
normal
throughout
the
rest
of
the
study.
Animals
feed
300
ppm
exhibited
roughed
fur
and
muscular
weakness
as
well
as
signs
of
cyanosis
after
9
days
of
treatment.
Animals
fed
1000
ppm
exhibited
similar
effects
after
6
days
and
those
receiving
3000
ppm
or
greater
doses
exhibited
effects
after
4
days.

All
mice
fed
APFO
lost
weight.
Reductions
in
body
weight
gain
were
followed
by
weight
losses
in
mice
fed
30,
100,
or
300
ppm.
A
dose­
related
pattern
was
seen
in
the
depressed
body
weights.

Relative
and
absolute
liver
weights
were
increased
in
mice
fed
30
ppm
or
more
APFO.
Gross
pathological
examination
of
kidneys
or
other
organs
besides
livers
is
not
discussed.
Treatmentrelated
changes
were
observed
in
the
livers
among
all
APFO
treated
animals
including
enlargement
and/
or
discoloration
of
1
or
more
liver
lobes.
Histopathologic
examination
of
all
APFO
treated
mice
revealed
diffuse
cytoplasmic
enlargement
of
hepatocytes
throughout
the
liver
(
pan
lobular
hypertrophy)
accompanied
by
focal
to
multifocal
cytoplasmic
vacuoles.
Degeneration
and
/
or
necrosis
of
hepatocytes
and
focal
bile
duct
proliferation
were
also
noted
in
mice
within
all
groups
(
Christopher
and
Marias,
1977).

Three
90­
day
subchronic
toxicity
studies
have
been
conducted.
One
was
conducted
in
rats
(
Goldenthal,
1978a),
one
was
conducted
in
rhesus
monkeys
(
Goldenthal,
1978b)
and
the
third
was
conducted
in
male
rats
(
Palazzolo,
1993).

In
the
monkey
study,
Goldenthal
(
1978b)
administered
rhesus
monkeys
(
2/
sex/
group)
doses
of
0,
3,
10,
30
or
100
mg/
kg/
day
perfluorooctanoic
acid
(
FC­
143)
in
0.5%
Methocel7
by
gavage
for
7
days/
week
for
90
days.
All
doses
were
given
in
a
constant
volume;
individual
daily
doses
were
based
upon
the
weekly
body
weights.
Animals
were
observed
twice
daily
for
general
physical
appearance
and
behavior
and
pharmacotoxic
signs.
General
physical
examinations
were
performed
during
the
control
period
and
monthly
during
the
study
period.
Individual
body
weights
were
recorded
weekly.
Blood
and
urine
samples
were
collected
once
during
the
control
period
and
at
1
and
3
months
of
the
study
for
hematology,
clinical
chemistry
and
urinalysis.
Monkeys
were
fasted
overnight
prior
to
the
collection
of
blood
and
urine
samples.
Organs
and
tissues
from
animals
that
were
sacrificed
at
the
end
of
the
study
and
from
animals
that
died
during
the
treatment
period
were
weighed,
examined
for
gross
pathology
and
samples
taken
for
histopathology.
Histopathology
was
performed
on
the
following
organs
from
all
monkeys
in
the
control
and
treatment
groups:
adrenals,
aorta,
bone,
brain,
esophagus,
eyes,
gallbladder,
heart
(
with
coronary
vessels),
duodenum,
ileum,
jejunum,
cecum,
colon,
rectum,
kidneys,
liver,
lung,
skin,
mesenteric
lymph
node,
retropharyngeal
lymph
node,
mammary
gland,
nerve
(
with
muscle),
spleen,
pancreas,
prostate/
uterus,
rib
junction
(
bone
marrow),
salivary
gland,
lumbar
spinal
cord,
pituitary,
stomach,
testes/
ovaries,
thyroid,
parathyroid,
thymus,
trachea,
tonsil,
tongue,
urinary
bladder,
vagina,
identifying
tattoo,
and
any
tissues(
s)
with
lesions.

All
monkeys
in
the
100­
mg/
kg/
day
groups
died
during
the
study.
The
first
death
occurred
during
week
2;
all
animals
were
dead
by
week
5.
Signs
and
symptoms
which
first
appeared
during
47
week
1
included
anorexia,
frothy
emesis
which
was
sometimes
brown
in
color,
pale
face
and
gums,
swollen
face
and
eyes,
slight
to
severe
decreased
activity,
prostration
and
body
trembling.
Three
monkeys
from
the
30­
mg/
kg/
day
group
died
during
the
study;
one
male
died
during
week
7
and
the
two
females
died
during
weeks
12
and
13.
Beginning
in
week
4,
all
four
animals
showed
slight
to
moderate
and
sometimes­
severe
decreased
activity.
One
monkey
had
emesis
and
ataxia,
swollen
face,
eyes
and
vulva,
as
well
as
pallor
of
the
face
and
gums.
Beginning
in
week
6,
two
monkeys
had
black
stools
and
one
monkey
had
slight
to
moderate
dehydration
and
ptosis
of
the
eyelids.

No
monkeys
in
the
3
or
10
mg/
kg/
day
groups
died
during
the
study.
Animals
in
the
3­
mg/
kg/
day­
dose
group
occasionally
had
soft
stools
or
moderate
to
marked
diarrhea;
frothy
emesis
was
also
occasionally
noted
in
this
group.
One
monkey
in
the
10
mg/
kg/
day
group
was
anorexic
during
week
4,
had
a
pale
and
swollen
face
in
week
7
and
had
black
stools
for
several
days
in
week
12.
The
other
animals
in
the
10­
mg/
kg/
day
groups
did
not
show
any
unusual
signs
or
symptoms.

Changes
in
body
weight
were
similar
to
the
controls
for
animals
from
the
3
and
10
mg/
kg/
day
dose
groups.
Monkeys
from
the
30
and
100
mg/
kg/
day
groups
lost
body
weight
after
week
1.
At
the
end
of
the
study,
this
loss
was
statistically
significant
for
the
one
surviving
male
in
the
30­
mg/
kg/
day
group
(
2.30
kg
vs
3.78
kg
for
the
control).

Hematology
values
at
the
end
of
the
1
and
3
months
of
treatment
were
similar
for
the
control
and
the
3
and
10
mg/
kg/
day
groups.
At
30
mg/
kg/
day,
the
surviving
male
had
decreased
numbers
of
erythrocytes,
decreased
hemoglobin,
decreased
hematocrit,
and
increased
platelets.
Prothrombin
time
and
activated
prothrombin
time
were
also
increased.
These
increases
were
apparent
at
1
month
but
were
much
more
marked
at
three
months.

Following
one
month
of
treatment,
glucose
was
significantly
elevated
in
the
3­
mg/
kg/
day
group
(
117
vs
89
mg/
100
ml
in
the
control).
The
authors
of
the
report
attribute
this
to
a
single
high
value
for
male
#
7366
who
had
a
value
of
131.
The
other
three
monkeys
in
the
3­
mg/
kg/
day
groups
had
levels
of
112,
105,
and
120­
mg/
100
ml.
Glucose
levels
in
the
10
and
30
mg/
kg/
day
groups
were
104
and
122­
mg/
100
ml,
respectively,
after
one
month
of
treatment.
At
three
months
of
treatment,
glucose
levels
were
81,
96,
88,
and
66­
mg/
100
ml
in
the
control,
3,
10
and
30
mg/
kg/
day
groups
respectively.

There
was
a
decrease
in
alkaline
phosphatase
levels
in
the
30­
mg/
kg/
day
group
(
365
vs
597
IU/
l
in
the
control)
at
one
month,
which
persisted
in
the
one
surviving
male
(
360
vs
851
IU/
l
in
the
control)
at
3
months.
Alkaline
phosphatase
levels
in
the
3­
and
10
mg/
kg/
day
groups
at
three
months
were
783
and
743
IU/
l
showing
a
dose­
related
trend
toward
decreased
levels.

SGOT
levels
were
reduced
in
the
30­
mg/
kg/
day
groups
at
one
month
(
59
vs
29
IU/
l
in
the
control)
and
in
the
one
surviving
male
at
3
months
(
88
vs
45
IU/
l
in
the
control).
SGPT
was
elevated
in
both
the
10
and
30
mg/
kg/
day
dose
groups
at
1
month;
the
levels
were
15,
34,
and
44
IU/
l
in
the
control,
10
and
30
mg/
kg/
day
groups,
respectively.
SGOT
levels
in
the
10­
mg/
kg/
day
48
group
were
comparable
to
the
controls
at
3
months
(
34
vs
31
IU/
l
in
the
control)
but
were
still
elevated
in
the
one
surviving
male
in
the
30­
mg/
kg/
day
dose
group
(
46
IU/
l).

Cholesterol
in
the
one
surviving
male
in
the
30
mg/
kg/
day
group
was
elevated
(
240
vs
165
mg/
100ml)
and
total
protein
and
albumin
in
this
animal
were
reduced.
Total
protein
was
5.52
vs
a
control
level
of
8.21
g/
100
ml
and
total
albumin
was
2.00
vs
a
control
level
of
4.82
g/
100
ml.

There
were
no
treatment
related
changes
in
urinalysis
studies
at
any
time
period
studied.

There
were
no
macroscopic
lesions
noted
at
gross
necropsy
of
any
animals
which
died
during
the
study
or
which
were
sacrificed
at
the
end
of
the
treatment
period.

The
following
changes
in
absolute
and
relative
organ
weight
changes
were
noted:
absolute
and
relative
weight
of
the
hearts
in
females
from
the
10
mg/
kg/
day
group
were
decreased;
absolute
brain
weight
of
females
from
this
same
group
were
also
decreased
and
relative
group
mean
weight
of
the
pituitary
in
males
from
the
3
mg/
kg/
day
group
was
increased.
The
significance
of
these
weight
changes
is
difficult
to
assess,
as
they
were
not
accompanied
by
morphologic
changes.

One
male
and
two
females
from
the
30
mg/
kg/
day
group
and
all
animals
from
the
100
mg/
kg/
day
group
had
marked
diffuse
lipid
depletion
in
the
adrenals.
All
males
and
females
from
the
30
and
100
mg/
kg/
day
groups
also
had
slight
to
moderate
hypocellularity
of
the
bone
marrow
and
moderate
atrophy
of
lymphoid
follicles
in
the
spleen.
One
female
from
the
30­
mg/
kg/
day
group
and
all
animals
in
the
100­
mg/
kg/
day
group
had
moderate
atrophy
of
the
lymphoid
follicles
in
the
lymph
nodes.
No
other
compound
related
lesions
were
seen
in
at
the
30
and
100
mg/
kg/
day
groups.
No
treatment
related
lesions
were
seen
in
the
organs
of
animals
from
the
3
and
10
mg/
kg/
day
groups.

The
levels
of
PFOA
in
the
serum
and
liver
are
presented
below
in
Table
3.
Individual
values
are
presented
so
there
are
double
entries
for
most
dose
levels.

Table
3.
Levels
of
PFOA
in
the
Serum
and
Liver
Dose
Serum
(
ppm)
Liver
(
ppm)
Liver
total
(
ug)

Males
Females
Males
Females
Males
Females
0
ND
1
0.05
0.07
3
5
3
53
65
3
7
250
350
3
48
50
ND
ND
ND
ND
10
45
79
9
ND
600
ND
10
71
71
ND
10
ND
750
30
ND
ND
125
80
8000
7500
30
145
ND
60
125
4000
9000
100
ND
ND
100
325
6000
20000
49
In
the
first
rat
study,
Goldenthal
(
1978a)
administered
CD
rats
(
5/
sex/
group)
dietary
levels
of
0,
10,
30,
100,
300,
and
1000
ppm
perfluorooctanoic
acid.
These
dose
levels
are
approximately
equivalent
to
0.6,
1.7,
5.6,
17.9,
and
63.5
mg/
kg/
day
in
males,
and
0.7,
2.3,
7.7,
22.4
and
76.5
mg/
kg/
day
in
females.
Animals
were
housed
individually
in
wire
mesh
cages
and
had
free
access
to
food
and
water.
Animals
were
observed
twice
daily
for
signs
of
toxicity
and
for
mortality.
Detailed
examinations
were
performed
once
a
week.
Individual
body
weight
and
food
consumption
were
recorded
weekly
during
the
pretest
and
treatment
periods.
Blood
and
urine
samples
were
collected
during
the
pretest
period
and
at
1
and
3
months
of
the
study
for
hematology
and
clinical
chemistry
and
urinalysis.
At
week
13,
the
serum
samples
were
frozen
and
shipped
to
the
sponsor
for
analysis.
Organs
and
tissues
from
animals
that
were
sacrificed
at
the
end
of
the
study
and
from
two
females
that
died
during
the
treatment
period
were
weighed,
examined
for
gross
pathology
and
samples
taken
for
histopathology.
Histopathology
was
performed
on
the
following
organs
from
rats
from
the
control,
100,
300,
and
1000
ppm
dose
groups:
brain
with
cervical
cord,
lumbar
spinal
cord,
peripheral
nerve,
eyes,
pituitary,
thyroid
with
parathyroid,
adrenals,
lung,
heart
with
coronary
vessels,
aorta,
spleen,
mesenteric
lymph
node,
thymus,
bone
with
marrow
(
sternum),
salivary
gland,
small
intestines
(
duodenum,
jejunum,
ileum)
colon,
pancreas,
liver,
kidneys,
urinary
bladder,
testes,
ovaries,
prostate,
uterus,
skin
(
mammary
gland),
any
tissue(
s)
with
gross
lesions.
Livers
from
rats
from
the
10
and
30­
ppm
dose
groups
were
also
examined
microscopically
and
liver
samples
from
all
dose
groups
were
frozen
and
sent
to
the
sponsor
for
analysis.

One
female
in
the
100
and
one
female
in
the
300­
ppm
group
died
during
collection
of
blood.
These
deaths
were
not
considered
to
be
treatment
related.
All
other
animals
survived
until
scheduled
sacrifice.

There
was
a
significant
reduction
in
mean
body
weight
in
males
in
the
1000­
ppm
group
(
362
g
vs
466
g
in
the
control
group).
Food
consumption
was
reduced
in
males
in
the
100,
300
and
1000­
ppm
groups,
but
the
differences
were
not
statistically
significant.

Males
in
the
30,
100,
300
and
1000­
ppm
groups
had
significantly
reduced
numbers
of
erythrocytes
at
the
end
of
the
treatment
period.
The
values
were
7.95,
7.05,
7.16,
6.72,
and
6.94
in
the
control,
30,
100,
300
and
1000­
ppm
groups,
respectively.
Males
had
reduced
leukocyte
values
compared
to
the
controls
in
all
dose
groups,
but
were
statistically
significant
at
the
300
ppm
group
only;
leukocyte
values
were
10.64,
8.88,
9.33,
9.35,
7.63,
and
8.06
in
the
control,
10,
30,
100,
300
and
1000
ppm
groups,
respectively.
A
similar
phenomenon
was
seen
with
hemoglobin
values,
which
were
reduced
at
all
dose
levels
but
were
significant
at
the
10­
ppm
dose
level
only.
Hemoglobin
values
were
16.2,
14.7,
15.0,
15.4,
14.9,
13.1
in
the
control,
10,
30,
100,
300
and
1000
ppm
groups,
respectively.
There
was
no
similar
effect
upon
the
hematological
parameters
of
female
rats
in
the
study.

Males
at
the
30,
100,
300,
and
1000­
ppm
dose
levels
had
increased
glucose
levels
(
mg/
100
ml),
which
were
statistically
significant
at
all
but
the
100­
ppm
dose
level.
Reported
glucose
levels
were
121,
120,
136,
134,
143
and
135
mg/
100
ml
for
the
0,
10,
30
100,
300
and
1000
ppm
50
groups,
respectively.
B.
U.
N.
levels
were
elevated
in
males
at
the
100,
300,
and
1000
ppm
dose
levels;
mean
values
at
90
days
were
20.4,
23.9
and
35.1
mg/
100
ml
for
the
three
dose
groups,
respectively,
compared
to
16.2
mg/
100
ml
for
the
controls.
Alkaline
phosphatase
was
elevated
in
males
in
the
100,
300,
and
1000­
ppm
groups;
the
levels
were
147,
204
and
212
IU/
l
for
the
three
groups,
respectively,
compared
to
104
IU/
l
for
the
controls.
Females
showed
no
similar
changes
in
biochemical
measurements.

Neither
males
nor
females
showed
any
treatment
related
changes
in
urinalysis
parameters
although
females
from
all
groups
showed
a
higher
frequency
of
occult
blood
in
the
urine
than
did
males.

The
only
gross
necropsy
observation
was
noted
in
males
at
the
1000­
ppm
dose
level.
These
animals
had
enlarged
livers
that
showed
varying
degrees
of
surface
discoloration.
Neither
females
from
the
1000­
ppm
dose
level
nor
males
or
females
from
the
lower
dose
levels
showed
such
effects.

Both
absolute
and
relative
liver
weights
were
significantly
increased
in
males
in
the
30,
300
and
1000­
ppm
groups
and
in
one
female
in
the
1000­
ppm
group.
Compound­
related
liver
lesions
occurred
in
all
male
rats
in
the
100,
300
and
1000­
ppm
groups.
These
lesions
consisted
of
focal
to
multifocal,
very
slight­
to­
slight
hypertrophy
of
hepatocytes
in
centrilobular
to
midzonal
regions
of
the
affected
liver
lobules.
In
some
instances
these
lesions
were
accompanied
by
an
increased
amount
of
yellowish­
brown
pigment
resembling
lipofuscin
in
the
cytoplasm
of
hepatocytes
and
occasionally
in
sinusoidal
lining
cells.
The
incidence
and
severity
of
the
lesions
was
more
pronounced
among
male
rats
at
the
1000­
ppm
dietary
level.

A
comparison
of
the
serum
levels
of
PFOA
is
shown
below
in
Table
4.
The
greater
toxicity
observed
in
the
males
than
in
the
females
is
due
to
the
gender
difference
in
elimination
as
demonstrated
by
the
differences
in
serum
PFOA
levels.

Table
4.
Comparison
of
Male
and
Female
PFOA
Serum
Levels
Dose
PFOA
in
Serum
(
ppm)
Males
Females
0
0
0
10
21
ND
30
34
0.15
100
36
ND
300
38
0.25
1000
49
0.65
ND
=
Not
Determined.

In
the
second
rat
study,
Palazzolo
(
1993)
administered
45­
55
male
Sprague­
Dawley
rats
per
group,
doses
of
1,
10,
30,
or
100
ppm
(
approximate
mean
compound
consumption
at
week
13
of
51
0.05,
0.47,
1.44,
and
4.97
mg/
kg/
day)
APFO
ad
libitum
in
the
diet
for
13
weeks.
Two
control
groups
(
a
nonpair­
fed
control
group
and
a
control
group
pair­
fed
to
the
100
ppm
dose
group)
were
also
exposed
during
that
period.
Following
the
13­
week
exposure
period,
10
animals
per
group
were
fed
basal
diet
for
an
additional
8­
weeks
post­
treatment
and
observed
for
any
signs
of
recovery.
All
test
diets
were
assayed
and
evaluated
for
test
material
homogeneity
and
stability.
All
animals
were
observed
twice
daily
for
mortality,
moribundity,
and
general
clinical
signs
of
toxicity.
Body
weights
were
recorded
once
before
exposures
began,
weekly
during
the
treatment
period,
and
then
on
the
day
of
necropsy.
Food
consumption
was
recorded
weekly
for
all
groups,
including
the
nonpair­
fed
control
group;
daily
for
the
pair­
fed
animals,
and
then
weekly
for
all
of
the
animals
retained
for
the
recovery
phase
of
the
study.
A
total
of
15
animals
per
group
were
sacrificed
following
4,
7,
or
13
weeks
of
treatment;
10
animals
per
group
were
sacrificed
after
13
weeks
of
treatment
and
following
8
weeks
of
non­
treatment.
Serum
samples
collected
from
10
animals
per
group
at
each
scheduled
sacrifice
during
treatment
and
from
5
animals
per
group
during
recovery
were
analyzed
for
estradiol,
total
testosterone,
luteinizing
hormones,
and
for
test
material
residue.
The
level
of
palmitoyl
CoA
oxidase,
an
indicator
of
peroxisome
proliferation,
was
analyzed
from
a
section
of
liver
that
was
obtained
from
5
animals
per
group
at
each
scheduled
sacrifice.
The
following
organs
from
all
animals
at
each
scheduled
sacrifice
were
weighed:
brain,
liver,
lungs,
testis
(
one),
seminal
vesicle,
prostate,
coagulating
gland,
and
urethra.
The
following
tissues
in
these
same
animals
were
preserved
in
10%
phosphate­
buffered
formalin
and
examined
macroscopically:
external
surface
of
the
body,
all
orifices,
the
cranial
cavity,
the
external
surfaces
of
the
brain
and
spinal
cord,
the
nasal
cavity
and
paranasal
sinuses;
the
thoracic,
abdominal,
and
pelvic
cavities
and
viscera;
and
also
examined
microscopically:
any
observed
lesions,
brain,
liver,
lungs,
testes
(
one),
seminal
vesicle,
prostate,
coagulating
gland,
and
urethra.
In
addition,
the
following
tissues
were
preserved
in
glutaraldehyde
for
electron
microscopic
examination:
brain,
liver,
lungs,
testes
(
one),
seminal
vesicle,
and
prostate.

In
the
analysis
of
the
data,
animals
in
groups
exposed
to
1,
10,
30,
and
100
ppm
APFO
were
compared
to
the
control
animals
in
the
nonpair­
fed
group,
while
the
data
from
the
pair­
fed
control
animals
were
compared
to
animals
exposed
to100
ppm
APFO.
All
test
diets
were
considered
to
be
homogeneous
and
stable
under
the
experimental
conditions.
All
animals
survived
to
scheduled
sacrifice,
with
the
exception
of
one
animal
in
the
100­
ppm
dosed­
group
that
was
sacrificed
on
week
4
due
to
severe
neck
sores
unrelated
to
treatment.
Twice­
daily
examinations
of
all
animals
were
unremarkable.
At
100
ppm,
significant
reductions
in
body
weights
were
seen
compared
to
the
pair­
fed
control
group
during
week
1
and
the
nonpair­
fed
control
group
during
weeks
1­
13
(
i.
e.,
throughout
treatment).
During
recovery,
however,
no
reductions
in
body
weights
were
apparent.
Body
weight
data
in
the
other
dosed­
groups
were
comparable
to
controls.
At
100
ppm,
mean
body
weight
gains
were
significantly
higher
than
the
pair­
fed
control
group
during
week
1
and
significantly
lower
than
the
nonpair­
fed
control
group
during
weeks
1­
13.
At
10
and
30
ppm,
mean
body
weight
gains
were
significantly
lower
than
the
nonpair­
fed
control
group
at
week
2.
These
differences
in
body
weight
gains
were
not
observed
during
the
recovery
period.
Significant
differences
in
food
consumption
were
observed
at
100
ppm
during
weeks
1
and
2
only,
when
compared
to
the
nonpair­
fed
control
group;
no
other
significant
differences
in
food
consumption
were
noted.
There
were
no
significant
differences
among
the
groups
for
any
of
the
hormones
evaluated
in
the
serum.
Likewise,
serum
52
analysis
of
test
material
residue
showed
no
increase
in
serum
APFO
levels
over
the
course
of
treatment.
Statistically
significant
higher
hepatic
palmitoyl
CoA
oxidase
activity
was
observed
at
30
and
100
ppm;
however,
this
effect
returned
to
control
levels
by
the
end
of
the
recovery
period.
At
10
ppm,
statistically
significant
higher
levels
of
hepatic
palmitoyl
CoA
oxidase
activity
were
observed
at
week
5
only.
Mean
enzyme
activities
were
highest
during
week
8
for
animals
exposed
to
10,
30,
and
100
ppm.
Significant
increases
in
absolute
and
relative
liver
weights
and
hepatocellular
hypertrophy
were
observed
at
weeks
4,
7,
or
13
in
the
10,
30
and
1000
ppm
groups.
The
authors
suggested
that
these
changes
might
be
associated
with
peroxisome
proliferation,
especially
since
increases
in
hepatic
palmitoyl
CoA
oxidase
activity
were
also
observed
at
this
dose
level
during
treatment.
During
recovery,
however,
none
of
the
liver
effects
were
observed,
indicating
that
these
treatment­
related
liver
effects
were
reversible.

Therefore,
under
the
conditions
of
this
study,
a
NOAEL
of
1.0
ppm
(
0.05
mg/
kg/
day)
and
a
LOAEL
of
10
ppm
(
0.47
mg/
kg/
day)
are
indicated
based
on
reductions
in
body
weight
and
body
weight
gain,
and
on
increases
in
absolute
and
relative
liver
weights
with
hepatocellular
hypertrophy.

Thomford
(
2001a,
b)
recently
conducted
a
range­
finding
and
a
6­
month
toxicity
study
in
cynomolgus
monkeys.
In
the
range­
finding
study,
Thomford
(
2001a)
administered
male
cynomolgus
monkeys
an
oral
capsule
containing
0,
2,
or
20
mg/
kg/
day
APFO
for
4
weeks.
There
were
3
monkeys
in
the
2
and
20
mg/
kg/
day
groups
and
one
monkey
in
the
control
group.
The
monkeys
weighed
2.1
to
3.6
kg
at
the
start
of
treatment.
Animals
were
observed
twice
daily
for
mortality
and
moribundity
and
were
examined
at
least
once
daily
for
signs
of
poor
health
or
abnormal
behavior.
Food
consumption
was
assessed
qualitatively.
Body
weight
data
were
recorded
before
the
start
of
treatment,
on
day
1
of
treatment
and
weekly
thereafter.
The
monkeys
were
fasted
overnight
and
blood
samples
were
collected
one
week
prior
to
the
start
of
the
study
and
on
day
30
for
clinical
hematology
and
clinical
chemistry
analyses,
and
hormone
and
PFOA
level.
Blood
for
clinical
chemistry
was
also
collected
from
each
animal
on
day
2
(
approximately
24
hours
after
the
first
dose).
Samples
were
analyzed
for
estradiol,
estrone,
estriol,
thyroid
stimulating
hormone,
total
and
free
triiodothyronine,
and
total
and
free
thyroxin.

At
scheduled
necropsy,
samples
of
the
right
lateral
lobe
of
the
liver
were
collected
from
each
animal
and
analyzed
for
palmitoyl
CoA
oxidase
activity.
Representative
samples
of
liver,
right
and
left
testes,
and
pancreas
were
collected
from
each
animal
for
cell
proliferation
evaluation
using
proliferation
cell
nuclear
antigen.
Bile
was
collected
from
each
animal
for
bile
acid
determination.
A
sample
of
liver
was
collected
from
each
animal
for
PFOA
concentration
analysis.
The
following
tissues
(
when
present)
or
representative
samples
were
collected
and
preserved
in
10
%
neutral­
buffered
formalin,
unless
otherwise
specified,
for
possible
future
examination:
adrenal
(
2),
aorta,
brain,
cecum,
colon,
duodenum,
epididymis
(
2),
esophagus,
eyes
[
preserved
in
Davidson's
fixative
(
2)],
femur
with
one
marrow
(
articular
surface
of
the
distal
end),
gallbladder,
heart,
ileum,
jejunum,
kidney
(
2),
lesions,
liver,
lung,
lymph
node
(
mesenteric)
mammary
gland,
muscle
(
thigh),
pancreas,
pituitary,
prostate,
rectum,
salivary
gland
[
mandibular
(
2)],
sciatic
nerve,
seminal
vesicle
(
2)
skin,
spinal
cord
(
cervical,
thoracic,
and
lumbar),
spleen,
sternum
with
bone
marrow,
stomach,
testis
(
2),
thymus,
thyroid
(
2)
with
parathyroid,
trachea,
53
and
urinary
bladder.
The
adrenals,
liver,
pancreas,
spleen,
and
testes
from
each
animal
were
embedded
in
paraffin,
sectioned,
stained
with
hematoxylin
and
eosin
and
examined
microscopically.

All
animals
survived
to
scheduled
sacrifice.
There
were
no
clinical
signs
of
toxicity
in
the
treated
groups
and
there
was
no
effect
on
body
weight.
Low
or
no
food
consumption
was
observed
for
one
animal
given
20
mg/
kg/
day.
No
food
consumption
was
noted
for
this
animal
on
day
12
and
low
food
consumption
was
noted
on
days
5,
7,
11,
14
17,
and
24,
respectively.
For
this
animal,
decreased
food
consumption
is
in
all
likelihood
related
to
APFO
administration.
There
were
no
effects
on
estradiol,
estriol,
thyroid
stimulating
hormone,
total
and
free
triiodothyronine,
and
total
and
free
thyroxin.
Estrone
levels
were
notably
lower
for
males
given
2
and
20
mg/
kg/
day
APFO.
There
was
no
evidence
of
peroxisome
proliferation
or
cell
proliferation
in
the
liver,
testes
or
pancreas
of
treated
monkeys.
No
adverse
effects
were
noted
in
either
gross
or
clinical
pathology
studies.

In
the
26­
week
study,
male
cynomolgus
monkeys
were
administered
APFO
by
oral
capsule
at
doses
of
0,
3,
10
or
30
mg/
kg/
day
for
26
weeks
(
Thomford,
2001b;
Butenhoff
et
al.,
2002).
At
study
initiation
the
monkeys
weighed
3.2
to
4.5
kg.
There
were
4
monkeys
in
the
3
mg/
kg/
day
group
and
6
monkeys
in
each
of
the
other
groups.
Dosing
of
animals
in
the
30
mg/
kg/
day
dose
group
was
stopped
from
days
11
 
21
because
of
toxicity.
When
dosing
was
resumed
on
day
22,
animals
received
20
mg/
kg/
day
and
this
group
was
designated
the
30/
20
mg/
kg/
day
group.
At
the
end
of
the
26­
week
treatment
period,
2
animals
in
the
control
and
10
mg/
kg/
day
groups
were
observed
for
a
13­
week
recovery
period.

Animals
were
observed
twice
daily
for
mortality
and
moribundity
and
were
examined
at
least
once
daily
for
signs
of
poor
health
or
abnormal
behavior;
food
consumption
was
assessed
qualitatively.
Ophthalmic
examinations
were
done
before
initiation
of
treatment
and
during
weeks
26
and
40.
Body
weight
data
were
recorded
weekly
before
the
start
of
treatment,
on
day
1
of
treatment
and
weekly
thereafter.
Blood
and
urine
samples
were
collected
for
clinical
hematology,
clinical
chemistry,
and
urinalysis
before
the
start
of
treatment
and
on
days
11,
31,
63,
91,
182,
217,
245
and
275.
Blood
samples
were
also
taken
for
hormone
determinations;
samples
were
analyzed
for
estradiol,
estrone,
estriol,
thyroid
stimulating
hormone,
total
and
free
triiodothyronine,
total
and
free
thyroxin,
and
testosterone.
Blood,
urine
and
feces
were
collected
during
week
2
and
every
2
weeks
thereafter
during
treatment
and
recovery
for
PFOA
concentration
analyses.
The
animals
were
not
fasted
before
collections.

At
scheduled
necropsy,
liver
samples
were
taken
for
determination
of
PFOA
levels.
The
right
lateral
lobe
of
the
liver
was
collected
from
each
animal
for
palmitoyl
CoA
oxidase
activity
analyses,
and
representative
samples
of
liver,
right
and
left
testes,
and
pancreas
were
collected
from
each
animal
for
cell
proliferation
evaluation
using
proliferation
cell
nuclear
antigen.
All
available
bile
was
collected
for
bile
acid
determination.
The
following
organs
were
weighed
at
scheduled
and
unscheduled
sacrifices;
paired
organs
were
weighed
separately:
adrenal
(
2),
brain,
epididymis
(
2),
kidney
(
2),
liver,
pancreas,
testis
(
2),
and
thyroid
(
2)
with
parathyroid.
Organ
to
body
weight
percentages
and
organ
to
brain
weight
ratios
were
calculated.
The
following
tissues
54
were
collected
for
histopathology:
adrenal
(
2),
aorta,
brain,
cecum,
colon,
duodenum,
epididymis
(
2),
esophagus,
eyes
[
preserved
in
Davidson's
fixative
(
2)],
femur
with
bone
marrow
(
articular
surface
of
the
distal
end),
gallbladder,
heart,
ileum,
jejunum,
kidneys
(
2),
lesions,
liver,
lung,
mesenteric
lymph
node,
mammary
gland,
pancreas,
pituitary,
prostate,
rectum,
salivary
gland
[
mandibular
(
2)],
sciatic
nerve,
seminal
vesicle
(
2),
skeletal
muscle
(
thigh),
skin,
spinal
cord
(
cervical,
thoracic,
and
lumbar),
spleen,
sternum
with
bone
marrow,
stomach,
testis
[(
2)
preserved
in
Bouin's
solution,
thymus,
thyroid
(
2)
with
parathyroid,
trachea
and
urinary
bladder.

Two
animals,
one
male
from
the
30/
20
mg/
kg/
day
dose
group
and
one
male
from
the
3
mg/
kg/
day
dose
group,
were
sacrificed
in
moribund
condition
during
the
study.
The
male
in
the
30/
20
mg/
kg/
day
dose
group
was
sacrificed
on
day
29.
This
animal
exhibited
signs
of
hypoactivity,
weight
loss,
few
or
no
feces,
low
or
no
food
consumption
and
the
entire
body
was
cold
to
the
touch
before
death.
Necropsy
revealed
esophageal
and
gastric
lesions
that
were
indicative
of
an
injury
that
occurred
during
dosing
and
liver
lesions
that
were
presumed
to
be
treatment
related.
The
animal
from
the
3
mg/
kg/
day
dose
group
was
sacrificed
on
day
137.
This
animal
showed
clinical
signs
of
limited
use
and
paralysis
of
the
hind
limbs,
ataxia
and
hypoactive
behavior,
few
feces
and
no
food
consumption.
The
cause
of
death
was
not
determined,
but
APFO
treatment
could
not
be
ruled
out.

Males
given
30
mg/
kg/
day
from
days
1­
11
had
clinical
signs
of
few
feces
and
low
food
consumption
and
they
lost
weight
during
week
1
of
treatment.
Based
on
these
signs,
treatment
was
stopped
on
day
11
and
was
not
resumed
until
day
22.
When
treatment
was
resumed,
the
dose
was
lowered
to
20
mg/
kg/
day;
this
group
was
then
designated
the
30/
20
mg/
kg/
day
group.
Of
the
remaining
animals
in
this
group,
only
2
tolerated
this
dose
level
for
the
remaining
23
weeks
of
treatment.
Treatment
of
three
males
given
30/
20
mg/
kg/
day
was
halted
on
days
43
(
week
7),
66
(
week
10),
and
81
(
week
12)
respectively.
Clinical
signs
in
these
animals
included
thin
appearance,
few
or
no
feces,
low
or
no
food
consumption,
and
weight
loss.
The
animals
appeared
to
recover
from
compound­
related
effects
within
3
weeks
after
cessation
of
treatment.

Mean
body
weight
changes
were
notably
lower
during
weeks
1
and
2
for
males
receiving
30
mg/
kg/
day.
During
week
2,
this
change
was
statistically
significant.
Treatment
was
stopped
on
day
11
and
when
it
was
resumed
at
20
mg/
kg/
day
on
day
21,
mean
body
weight
changes
were
significantly
lower
than
controls
during
weeks
7,
9
and
24.
Overall
mean
body
weight
changes
through
week
27
were
notably
lower
for
the
males
in
the
30/
20
mg/
kg/
day
group.
There
was
an
increased
incidence
of
low
or
no
food
consumption
for
animals
in
the
30/
20
mg/
kg/
day
group
that
was
considered
to
be
treatment
related.

There
were
no
consistent
or
clearly
dose­
related
effects
on
estrone,
estradiol,
estriol,
thyroid
stimulating
hormone,
or
testosterone
that
were
seen
in
any
treatment
group
over
time.
In
general,
thyroid
hormones
were
decreased
beginning
on
day
35
in
animals
in
the
10
or
30/
20
mg/
kg/
day
groups.
Triiodothyronine
remained
depressed
through
day
183
in
the
30/
20
mg/
kg
group.
Total
thyroxin
was
decreased
beginning
on
day
35
in
animals
administered
10
or
30/
20
mg/
kg
APFO.
The
effect
on
thyroxin
was
most
pronounced
at
day
35
in
animals
administered
10
mg/
kg/
day
and
day
66
in
animals
administered
30/
20
mg/
kg/
day.
Thereafter,
the
effect
began
55
to
diminish
and
recovery
was
observed
either
in
the
last
3
months
of
dosing
or
during
the
recovery
phase
of
the
experiment.
No
changes
in
cholecystokinin
concentrations
were
seen
over
time
in
either
dose
group.

Two
males
from
the
30/
20
mg/
kg/
day
group,
the
one
sacrificed
on
day
29
and
one
for
whom
treatment
was
stopped
because
of
poor
health,
had
moderate
to
marked
increased
serum
enzyme
concentrations
(
i.
e.
aspartate
aminotransferase,
alanine
aminotransferase,
sorbitol
dehydrogenase,
and
creatine
kinase)
and
a
mildly
increased
serum
bile
acid
concentration.
At
30/
20
mg/
kg/
day,
there
were
slight
increases
in
triglyceride
concentrations
and
mild
to
moderate
decreases
in
absolute
neutrophil
count,
total
protein
concentration
and
albumin
concentration.
Two
animals
in
the
30/
20
mg/
kg/
day
group
had
markedly
increased
serum
enzyme
activities
and
mildly
increased
serum
bile
acid
concentrations.
At
3
mg/
kg/
day
or
10
mg/
kg/
day,
APFO
had
no
effects
on
hematology,
coagulation,
clinical
chemistry,
or
urinalysis.
There
was
no
evidence
of
persistent
or
delayed
toxic
effects
on
clinical
pathology
test
results
during
the
recovery
period.

At
terminal
sacrifice
at
26
weeks,
there
were
statistically
significant
increases
in
absolute
and
relative
liver
weights
in
all
dose
groups.
The
absolute
liver
weights
were
increased
by
35,
38
and
50%
in
the
3,
10
and
30/
20
mg/
kg/
day
groups,
respectively;
relative
liver
weights
were
increased
by
20,
20
and
60%
in
the
3,
10
and
30/
20
mg/
kg/
day
groups,
respectively.
In
addition,
mean
liver­
to­
brain
weight
was
significantly
increased
in
the
10
mg/
kg/
day
group.
The
increased
liver
weights
were
considered
treatment
related.
There
were
no
treatment­
related
macroscopic
or
microscopic
changes
in
any
organs
at
the
terminal
sacrifice,
including
liver,
adrenal,
spleen,
pancreas,
and
testis.
While
there
was
no
evidence
of
enhanced
cell
proliferation
in
the
testes
or
pancreas
of
the
treated
monkeys,
the
findings
in
the
liver
were
equivocal.
There
was
evidence
of
mitochondrial
proliferation
in
the
livers
of
treated
monkeys
(
Butenhoff
et
al.,
2002).

At
the
recovery
sacrifice,
there
were
no
treatment­
related
effects
on
terminal
body
weights
or
on
absolute
or
relative
organ
weights
at
recovery
sacrifice
indicating
that
the
liver
weight
changes
seen
at
terminal
sacrifice
were
reversible
over
time.
There
were
no
treatment­
related
macroscopic
or
microscopic
changes
at
the
recovery
sacrifice.

The
results
of
the
analyses
of
PFOA
in
the
serum,
liver,
urine
and
feces
are
presented
in
Tables
5­
8,
respectively
(
3M
Environmental
Laboratory,
2001c).
Low
levels
of
PFOA
were
often
detected
in
the
sera,
liver,
urine
and
feces
of
the
control
animals.
The
levels
in
treated
animals
were
significantly
higher
than
those
seen
in
the
controls.
In
general,
PFOA
levels
in
the
sera
of
test
animals
increased
with
dose
but
decreased
over
time
during
treatment.
On
day
9
of
treatment,
the
serum
levels
were
126
±
36.1
µ
g/
mL
in
the
3
mg/
kg/
day
group
and
1597
±
2392
µ
g/
mL
in
the
30/
20
mg/
kg/
day
group,
and
during
weeks
26/
27,
the
serum
levels
were
52.5
±
9.14
µ
g/
mL
in
the
3
mg/
kg/
day
group
and
51.5
±
77.6
µ
g/
mL
in
the
30/
20
mg/
kg/
day
group.
PFOA
levels
in
serum
decreased
over
time
during
recovery
until
they
reached
1.18
±
0.827
µ
g/
mL
in
the
10
mg/
kg/
day
group
as
compared
to
0.0738
±
0.00256
in
the
controls
during
week
40.

A
similar
trend
was
seen
in
the
urine.
On
day
9,
the
levels
in
the
urine
were
73.5
±
38.1,
221
±
56
124,
and
909
±
209
µ
g/
mL
in
the
3,
10
and
30/
20
mg/
kg/
day
groups,
respectively,
and
during
week
26,
the
levels
were
51.6
±
13.7,
109
±
75.2,
and
19.2
±
27.0
µ
g/
g,
respectively,
in
these
same
groups.
During
week
2,
the
levels
in
the
feces
were
less
than
the
limit
of
quantitation
in
the
control
animals,
and
7.43
±
6.54,
15.4
±
10.2
and
56.6
±
73.7
ug/
g
in
the
3,
10,
and
30/
20
mg/
kg/
day
groups,
respectively;
during
week
26
the
levels
were
2.92
±
1.35,
43.0
±
36.9
and
10.3
±
20.8
µ
g/
g
in
the
3,
10,
and
30/
20
mg/
kg/
day
groups,
respectively.
There
is
no
explanation
for
the
high
levels
of
PFOA
seen
in
the
feces
of
the
control
animals
during
week
22.
During
the
recovery
period,
PFOA
levels
in
both
urine
and
feces
fell
to
levels
that
were
comparable
to
control
levels.

Under
the
conditions
of
the
study,
the
LOAEL
was
3
mg/
kg/
day
(
liver
toxicity
and
possibly
mortality)
and
a
NOAEL
was
not
established.
57
Table
5
Average
PFOA
Concentrations
in
the
Serum
(
µ
g/
mL)
of
Treated
Animals
0
mg/
kg/
day
3
mg/
kg/
day
10
mg/
kg/
day
30/
20
mg/
kg/
day
Time
Point
Day
9
0.0613
±
0.0472
126
±
36.1
189
±
48.9
1597
±
2392
Week
4
0.0206
±
0.0105
98.4
±
42.7
172
±
71.9
1084
±
1839
Week
6
0.103
±
0.0113
102
±
33.6
95.8
±
20.6
145
±
21.6
Week
8
<
LOQ
94.7
±
27.3
97.6
±
23.5
166
±
98.9
Week
10
0.126
±
0.0348
105
±
37.7
93.6
±
13.7
237
±
158
Week
12
0.123
±
0.0507
79.6
±
25.9
90.6
±
24.5
140
±
77.5
Week
14
0.162
±
0.0643
90.0
±
28.9
92.2
±
27.6
79.4
±
28.3
Week
16
0.128
±
0.0721
68.6
±
26.0
98.5
±
42.3
83.9
±
58.5
Week
18
0.183
±
0.0637
29.4
±
25.4
17.4
±
14.3
36.2
±
21.2
Week
20
0.224
±
0.0730
33.0
±
26.0
96.6
±
26.6
97.7
±
129
Week
22
0.232
±
0.131
77.6
±
24.9
105
±
36.3
58.6
±
45.6
Week
24
<
LOQ
72.2
±
68.8
90.9
±
21.2
62.7
±
74.3
Week
26
0.209
±
0.156
118
±
27.5
77.4
±
16.9
77.8
±
126
Week
26/
27
0.223
±
0.105
52.5
±
9.14
74.1
±
33.1
51.5
±
77.6
Week
28
0.181
±
0.0391
NS
25.9
±
7.07
NS
Week
30
0.144
±
0.0238
NS
11.3
±
5.11
NS
Week
32
0.110
±
0.0216
NS
7.60
±
3.90
NS
Week
34
0.0861
±
0.0256
NS
3.97
±
2.09
NS
Week
36
0.111
±
0.0445
NS
2.75
±
1.88
NS
Week
38
0.0941
±
0.0324
NS
1.84
±
1.40
NS
Week
40
0.0738
±
0.00256
NS
1.18
±
0.827
NS
LOQ
=
Limit
of
Quantitation;
NS
=
No
Sample;
Results
are
expressed
as
group
averages
±
the
standard
deviation
associated
with
that
group.
Data
are
accurate
to
within
one
SD
of
the
average
fortified
sample
recovery.
The
average
fortified
sample
recovery
was
94%
with
an
SD
of
11%.
58
Table
6
Average
PFOA
Concentrations
in
the
Liver
(
µ
g/
g)
of
Treated
Animals
0
mg/
kg/
day
3
mg/
kg/
day
10
mg/
kg/
day
30/
20
mg/
kg/
day
Time
Point
Week
20
NS
18.3
NS
NS
Week
27
0.117
±
0.0730
15.3
±
3.02
14.0
±
7.55
42.8
±
63.3
Week
40
<
LOQ
NS
0.114
±
0.0441
NS
LOQ
=
Limit
of
Quantitation;
NS
=
No
Sample
Results
are
expressed
as
group
averages
±
the
standard
deviation
associated
with
that
group.
Data
are
accurate
to
within
one
SD
of
the
average
fortified
sample
recovery.
The
average
fortified
sample
recovery
was
90%
with
an
SD
of
26%.
59
Table
7.
Average
PFOA
Concentrations
in
the
Urine
(
µ
g/
mL)
of
Treated
Animals
0
mg/
kg/
day
3
mg/
kg/
day
10
mg/
kg/
day
30/
20
mg/
kg/
day
Time
Point
Week
2
<
LOQ
73.5
±
38.1
221
±
124
909
±
269
Week
4
0.152
±
0.337
54.9
±
4.62
190
±
91.6
240
±
161
Week
6
0.0587
±
0.0716
65.7
±
46.9
128
±
50.0
272
±
140
Week
8
0.0161
±
0.00940
47.6
±
20.6
206
±
73.1
180
±
109
Week
10
0.0177
±
0.00114
39.9
±
18.7
175
±
92.3
359
±
449
Week
12
0.0141
±
0.00648
48.8
±
18.8
201
±
92.7
118
±
111
Week
14
7.96
±
19.5
63.1
±
47.2
139
±
52.5
72.9
±
84.0
Week
16
0.0299
±
0.0339
50.2
±
21.0
139
±
57.0
50.2
±
67.9
Week
18
0.0256
±
0.0248
37.7
±
19.3
186
±
63.9
43.1
84.6
Week
20
0.0211
±
0.0130
52.1
±
9.63
144
±
135
44.0
±
59.9
Week
22
0.0231
±
0.00688
95.8
±
80.8
158
±
78.4
98.5
±
134
Week
24
0.0125
±
0.00749
46.3
±
8.52
157
±
63.3
56.0
±
77.2
Week
26
0.0268
±
00.0265
51.6
±
13.7
109
±
75.2
19.2
±
27.0
Week
28
0.118
±
0.142
NS
0.327
±
0.0182
NS
Week
30
<
LOQ
NS
0.361
±
0.118
NS
Week
32
<
LOQ
NS
0.114
±
0.0608
NS
Week
34
<
LOQ
NS
0.121
±
0.0305
NS
Week
36
0.0117
±
0.00841
NS
0.0502
±
0.0166
NS
Week
38
<
LOQ
NS
0.0284
±
0.0102
NS
Week
40
<
LOQ
NS
0.0254
±
0.0101
NS
LOQ
=
Limit
of
Quantitation;
NS
=
No
Sample;
Results
are
expressed
as
group
averages
±
the
standard
deviation
associated
with
that
group.
Data
are
accurate
to
within
one
SD
of
the
average
fortified
sample
recovery.
The
average
fortified
sample
recovery
was
88%
with
an
SD
of
17%.
60
Table
8
Average
PFOA
Concentrations
in
the
Feces
(
µ
g/
g)
of
Treated
Animals
0
mg/
kg/
day
3
mg/
kg/
day
10
mg/
kg/
day
30/
20
mg/
kg/
day
Time
Point
Week
2
<
LOQ
7.43
±
6.54
15.4
±
10.2
56.6
±
73.7
Week
4
0.0214
±
0.0178
10.4
±
12.0
23.4
±
10.6
22.0
±
6.23
Week
6
0.108
±
0.00192
12.1
±
14.1
23.3
±
8.46
101
±
86.7
Week
8
0.0782
±
0.103
9.46
±
9.21
41.0
±
25.0
36.7
±
34.2
Week
10
<
LOQ
3.96
±
3.68
26.0
±
17.4
48.0
±
34.0
Week
12
0.0498
±
0.0894
7.15
±
5.65
10.3
±
6.07
32.0
±
42.9
Week
14
0.139
±
0.308
7.50
±
2.43
27.2
±
29.4
19.2
±
25.2
Week
16
0.0572
±
0.0762
6.88
±
2.62
31.4
±
23.3
18.2
±
28.8
Week
18
0.258
±
0.654
5.72
±
7.15
17.3
±
13.3
22.1
±
31.7
Week
20
0.405
±
1.08
6.81
±
4.89
52.4
±
39.5
37.8
±
58.1
Week
22
15.5
±
36.9
13.8
±
5.22
39.5
±
21.0
25.2
±
36.0
Week
24
0.517
±
1.13
6.22
±
5.45
40.5
±
21.8
34.6
±
47.7
Week
26
0.0172
±
0.00892
2.92
±
1.35
43.0
±
36.9
10.3
±
20.8
Weeks
28­
34
0.279
±
0.732
NS
0.387
±
0.372
NS
Weeks
36­
40
0.0103
±
0.00684
NS
0.0336
±
0.0313
NS
LOQ
=
Limit
of
Quantitation;
NS
=
No
Sample
Results
are
expressed
as
group
averages
±
the
standard
deviation
associated
with
that
group.
Data
are
accurate
to
within
one
SD
of
the
average
fortified
sample
recovery.
The
average
fortified
sample
recovery
was
117%
with
an
SD
of
22%.

3.6
Developmental
Toxicity
Studies
in
Animals
Three
prenatal
developmental
toxicity
studies
of
APFO
have
been
conducted,
one
inhalation
and
two
oral
studies.

The
first
of
these
studies
was
an
oral
developmental
toxicity
study
in
rats
(
Gortner,
1981).
Based
on
the
results
of
a
range­
finding
study,
an
upper
dose
level
of
150
mg/
kg/
day
was
set
for
the
definitive
study
in
which
five
groups
of
22
time­
mated
Sprague­
Dawley
rats
were
administered
0,
0.05,
1.5,
5,
and
150
mg/
kg/
day
APFO
in
distilled
water
by
gavage
on
gestation
days
(
GD)
6­
15.
Doses
were
adjusted
according
to
body
weight.
Dams
were
monitored
on
GD
3­
20
for
clinical
signs
of
toxicity.
Individual
body
weights
were
recorded
on
GD
3,
6,
9,
12,
15,
and
20.
61
Animals
were
sacrificed
on
GD
20
by
cervical
dislocation
and
the
ovaries,
uteri,
and
contents
were
examined
for
the
number
of
corpora
lutea,
number
of
viable
and
non­
viable
fetuses,
number
of
resorption
sites,
and
number
of
implantation
sites.
Fetuses
were
weighed
and
sexed
and
subjected
to
external
gross
necropsy.
Approximately
one­
third
of
the
fetuses
were
fixed
in
Bouin's
solution
and
examined
for
visceral
abnormalities
by
free­
hand
sectioning.
The
remaining
fetuses
were
subjected
to
skeletal
examination
using
alizarin
red.

Signs
of
maternal
toxicity
consisted
of
statistically
significant
reductions
in
mean
maternal
body
weights
on
GD
9,
12,
and
15
at
the
high­
dose
group
of
150
mg/
kg/
day.
Mean
maternal
body
weight
on
GD
20
continued
to
remain
lower
than
controls,
although
the
difference
was
not
statistically
significant.
Other
signs
of
maternal
toxicity
that
occurred
only
at
the
high­
dose
group
included
ataxia
and
death
in
three
rat
dams.
No
other
effects
were
reported.
Administration
of
APFO
during
gestation
did
not
appear
to
affect
the
ovaries
or
reproductive
tract
of
the
dams.
Under
the
conditions
of
the
study,
a
NOAEL
of
5
mg/
kg/
day
and
a
LOAEL
of
150
mg/
kg/
day
for
maternal
toxicity
were
indicated.

A
significantly
higher
incidence
in
fetuses
with
one
missing
sternebrae
was
observed
at
the
highdose
group
of
150
mg/
kg/
day;
however
this
skeletal
variation
also
occurred
in
the
controls
and
the
other
three
dose
groups
(
at
similar
incidence
but
lower
than
the
high­
dose
group)
and
therefore
was
not
considered
to
be
treatment­
related.
No
significant
differences
between
treated
and
control
groups
were
noted
for
other
developmental
parameters
that
included
the
mean
number
of
males
and
females,
total
and
dead
fetuses,
the
mean
number
of
resorption
sites,
implantation
sites,
corpora
lutea
and
mean
fetus
weights.
Likewise,
a
fetal
lens
finding
initially
described
as
a
variety
of
abnormal
morphological
changes
localized
to
the
area
of
the
embryonal
nucleus,
was
later
determined
to
be
an
artifact
of
the
free­
hand
sectioning
technique
and
therefore
not
considered
to
be
treatment­
related.
Under
the
conditions
of
the
study,
a
NOAEL
for
developmental
toxicity
of
150
mg/
kg/
day
(
highest
dose
group)
was
indicated.

A
second
oral
prenatal
developmental
toxicity
study
was
conducted
in
rabbits
(
Gortner,
1982).
Based
on
the
results
of
a
range­
finding
study,
an
upper
dose
level
of
50
mg/
kg/
day
was
set
for
the
definitive
study
in
which
four
groups
of
18
pregnant
New
Zealand
White
rabbits
were
administered
0,
1.5,
5,
and
50
mg/
kg/
day
APFO
in
distilled
water
by
gavage
on
gestation
days
(
GD)
6­
18.
Pregnancy
was
established
in
each
sexually
mature
female
by
i.
v.
injection
of
pituitary
lutenizing
hormone
in
order
to
induce
ovulation,
followed
by
artificial
insemination
with
0.5
ml
of
pooled
semen
collected
from
male
rabbits;
the
day
of
insemination
was
designated
as
day
0
of
gestation.
A
constant
dose
volume
of
1
ml/
kg
was
administered.
Individual
body
weights
were
measured
on
GD
3,
6,
9,
12,
15,
18,
and
29.
The
does
were
observed
daily
on
GD
3­
29
for
abnormal
clinical
signs.
On
GD
29,
the
does
were
euthanized
and
the
ovaries,
uterus
and
contents
examined
for
the
number
of
corpora
lutea,
live
and
dead
fetuses,
resorptions
and
implantation
sites.
Fetuses
were
examined
for
gross
abnormalities
and
placed
in
a
370
C
incubator
for
a
24­
hour
survival
check.
Pups
were
subsequently
euthanized
and
examined
for
visceral
and
skeletal
abnormalities.
A
blood
sample
was
taken
from
six
does
prior
to
dosing
and
then
on
GD
18
and
29;
a
liver
sample
was
taken
from
the
same
animals
on
GD
29.
All
samples
were
sent
to
the
sponsor
for
analysis.
This
information
was
unavailable
at
the
time
of
this
62
review.

Signs
of
maternal
toxicity
consisted
of
statistically
significant
transient
reductions
in
body
weight
gain
on
GD
6­
9
when
compared
to
controls;
body
weight
gains
returned
to
control
levels
on
GD
12­
29.
Administration
of
APFO
during
gestation
did
not
appear
to
affect
the
ovaries
or
reproductive
tract
contents
of
the
does.
No
clinical
or
other
treatment­
related
signs
were
reported.
Under
the
conditions
of
the
study,
a
NOAEL
of
50
mg/
kg/
day,
the
highest
dose
tested,
for
maternal
toxicity
was
indicated.

No
significant
differences
were
noted
between
controls
and
treated
groups
for
the
number
of
males
and
females,
dead
or
live
fetuses,
and
fetal
weights.
Likewise,
there
were
no
significant
differences
reported
for
the
number
of
resorption
and
implantation
sites,
corpora
lutea,
the
conception
incidence,
abortion
rate,
or
the
24­
hour
mortality
incidence
of
the
fetuses.
Gross
necropsy
and
skeletal/
visceral
examinations
were
unremarkable.
The
only
sign
of
developmental
toxicity
consisted
of
a
dose­
related
increase
in
a
skeletal
variation,
extra
ribs
or
13th
rib,
with
statistical
significance
at
the
high­
dose
group
(
38%
at
50
mg/
kg/
day,
30%
at
5
mg/
kg/
day,
20%
at
1.5
mg/
kg/
day,
and
16
%
at
0
mg/
kg/
day).
A
statistically
significant
increase
in
13th
ribs­
spurred
occurred
in
the
mid­
dose
group
of
5
mg/
kg/
day;
however,
the
biological
significance
of
this
effect
is
uncertain
since
in
both
the
high­
and
low­
dose
groups,
this
effect
occurred
at
the
same
rate
and
was
not
statistically
significantly
different
from
controls.
Therefore,
under
the
conditions
of
the
study,
a
LOAEL
for
developmental
toxicity
of
50
mg/
kg/
day
(
highest
dose
group)
was
indicated.

Staples
et
al.
(
1984)
also
conducted
a
developmental
toxicity
study
of
APFO.
The
study
design
consisted
of
an
inhalation
and
an
oral
portion,
each
with
two
trials
or
experiments.
The
first
trial
was
the
teratology
portion
of
the
study,
in
which
the
dams
were
sacrificed
on
GD
21;
while
in
the
second
trial,
the
dams
were
allowed
to
litter
and
the
pups
were
sacrificed
on
day
35­
post
partum.
For
the
inhalation
portion
of
the
study,
the
two
trials
consisted
of
12
pregnant
Sprague­
Dawley
rats
per
group
exposed
to
0,
0.1,
1,
10,
and
25
mg/
m3
APFO
for
6
hours/
day,
on
GD
6­
15.
In
the
oral
portion
of
the
study,
25
and
12
Sprague­
Dawley
rats
for
the
first
and
second
trials,
respectively,
were
administered
0
and
100
mg/
kg/
day
APFO
in
corn
oil
by
gavage
on
GD
6­
15.
For
both
routes
of
administration,
females
were
mated
on
an
as­
needed
basis
and
when
the
number
of
mated
females
was
bred,
they
were
ranked
within
breeding
days
by
body
weight
and
assigned
to
groups
by
rotation
in
order
of
rank.
Finally,
two
additional
groups
(
six
dams
per
group)
were
added
to
each
trial
that
was
pair­
fed
to
the
10
and
25
mg/
m3
groups.

For
the
teratology
portion
of
the
study
(
trial
one),
dams
were
weighed
on
GD
1,
6,
9,
13,
16,
and
21
and
observed
daily
for
abnormal
clinical
signs.
On
GD
21,
the
dams
were
sacrificed
by
cervical
dislocation
and
examined
for
any
gross
abnormalities,
liver
weights
were
recorded
and
the
reproductive
status
of
each
animal
was
evaluated.
The
ovaries,
uterus
and
contents
were
examined
for
the
number
of
corpora
lutea,
live
and
dead
fetuses,
resorptions
and
implantation
sites.
Pups
(
live
and
dead)
were
counted,
weighed
and
sexed
and
examined
for
external,
visceral,
and
skeletal
alterations.
The
heads
of
all
control
and
high­
dosed
group
fetuses
were
examined
for
visceral
alterations
as
well
as
macro­
and
microscopic
evaluation
of
the
eyes.
63
For
trial
two,
in
which
the
dams
were
allowed
to
litter,
the
procedure
was
the
same
as
that
for
trial
one
up
to
GD
21.
Two
days
before
the
expected
day
of
parturition,
each
dam
was
housed
in
an
individual
cage.
The
date
of
parturition
was
noted
and
designated
Day
1
PP.
Dams
were
weighed
and
examined
for
clinical
signs
on
Days
1,
7,
14,
and
22
PP.
On
Day
23
PP
all
dams
were
sacrificed.
Pups
were
counted,
weighed,
and
examined
for
external
alterations.
Each
pup
was
subsequently
weighed
and
inspected
for
adverse
clinical
signs
on
Days
4,
7,
14,
and
22
PP.
The
eyes
of
the
pups
were
also
examined
on
Days
15
and
17
PP
for
the
inhalation
portion
and
on
Days
27
and
31
PP
for
the
gavage
portion
of
the
study.
Pups
were
sacrificed
on
Day
35
PP
and
examined
for
visceral
and
skeletal
alterations.

Inhalation
Exposure
Trial
One:

Treatment­
related
clinical
signs
of
maternal
toxicity
for
trial
one
(
teratology)
occurred
at
10
and
25
mg/
m3
and
consisted
of
wet
abdomens,
chromodacryorrhea,
chromorhinorrhea,
a
general
unkempt
appearance,
and
lethargy
in
four
dams
at
the
end
of
the
exposure
period
(
highconcentration
group
only).
Three
out
of
12
dams
died
during
treatment
at
25
mg/
m3
(
on
GD
12,
13,
and
17).
Food
consumption
was
significantly
reduced
at
both
10
and
25
mg/
m3;
however,
no
significant
differences
were
noted
between
treated
and
pair­
fed
groups.
Significant
reductions
in
body
weight
were
also
observed
at
these
concentrations,
with
statistical
significance
at
the
highconcentration
only.
Likewise,
statistically
significant
increases
in
mean
liver
weights
were
seen
at
the
high­
concentration
group.
Under
the
conditions
of
the
study,
a
NOAEL
and
LOAEL
for
maternal
toxicity
of
1
and
10
mg/
m3,
respectively,
were
indicated.

No
effects
were
observed
on
the
maintenance
of
pregnancy
or
the
incidence
of
resorptions.
Mean
fetal
body
weights
were
significantly
decreased
in
the
25­
mg/
m3
groups
and
in
the
control
group
pair­
fed
25
mg/
m3.
A
detailed
microscopic
visceral
and
eye
examination
of
the
fetuses
did
not
reveal
any
treatment­
related
effects;
however
in
the
control
group
that
was
pair­
fed
25
mg/
m3,
a
statistically
significant
increased
incidence
of
fetuses
with
partially
ossified
sternebrae
was
observed.
Under
the
conditions
of
the
study,
a
NOAEL
and
LOAEL
for
developmental
toxicity
of
10
and
25
mg/
m3,
respectively,
were
indicated.

Trial
Two:

Clinical
signs
of
maternal
toxicity
seen
at
10
and
25
mg/
m3
were
similar
in
type
and
incidence
to
those
described
for
trial
one.
Maternal
body
weight
gain
during
treatment
at
25
mg/
m3
was
less
than
controls,
although
the
difference
was
not
statistically
significant.
In
addition,
2
out
of
12
dams
died
during
treatment
at
25
mg/
m3.
No
other
treatment­
related
effects
were
reported,
nor
were
any
adverse
effects
noted
for
any
of
the
measurements
of
reproductive
performance.
Under
the
conditions
of
the
study,
a
NOAEL
and
LOAEL
for
maternal
toxicity
of
1
and
10
mg/
m3,
respectively,
were
indicated.
64
Signs
of
developmental
toxicity
in
this
group
consisted
of
statistically
significant
reductions
in
pup
body
weight
on
Day
1
PP
(
6.1
g
at
25
mg/
m3
vs.
6.8
g
in
controls).
On
Days
4
and
22
PP,
pup
body
weights
continued
to
remain
lower
than
controls,
although
the
difference
was
not
statistically
significant
(
Day
4
PP:
9.7
g
at
25
mg/
m3
vs.
10.3
in
controls;
Day
22
PP:
49.0
g
at
25
mg/
m3
vs.
50.1
in
controls).
No
significant
effects
were
reported
following
external
examination
of
the
pups
or
with
ophthalmoscopic
examination
of
the
eyes.
Under
the
conditions
of
the
study,
a
NOAEL
and
LOAEL
for
developmental
toxicity
of
10
and
25
mg/
m3,
respectively,
were
indicated.

Oral
Exposure
Trial
One:

Three
out
of
25
dams
died
during
treatment
of
100
mg/
kg
APFO
during
gestation
(
one
death
on
GD
11;
two
on
GD
12).
Clinical
signs
of
maternal
toxicity
in
the
dams
that
died
were
similar
to
those
seen
with
inhalation
exposure.
Food
consumption
and
body
weights
were
reduced
in
treated
animals
compared
to
controls.
No
adverse
signs
of
toxicity
were
noted
for
any
of
the
reproductive
parameters
such
as
maintenance
of
pregnancy
or
incidence
of
resorptions.
Likewise,
no
significant
differences
between
treated
and
control
groups
were
noted
for
fetal
weights,
or
in
the
incidences
of
malformations
and
variations;
nor
were
there
any
effects
noted
following
microscopic
examination
of
the
eyes.

Trial
Two:

Similar
observations
for
clinical
signs
were
noted
for
the
dams
as
in
trial
one.
Likewise,
no
adverse
effects
on
reproductive
performance
or
in
any
of
the
fetal
observations
were
noted.

3.7
Reproductive
Toxicity
Studies
in
Animals
York
(
2002)
conducted
a
two­
generation
reproductive
toxicity
study
of
APFO.
Five
groups
of
30
Sprague­
Dawley
rats
per
sex
per
dose
group
were
administered
APFO
by
gavage
at
doses
of
0,
1,
3,
10,
and
30
mg/
kg/
day
six
weeks
prior
to
and
during
mating.
Treatment
of
the
F0
male
rats
continued
until
mating
was
confirmed,
and
treatment
of
the
F0
female
rats
continued
throughout
gestation,
parturition,
and
lactation.

The
F0
animals
were
examined
twice
daily
for
clinical
signs,
abortions,
premature
deliveries,
and
deaths.
Body
weights
of
F0
male
rats
were
recorded
weekly
during
the
dosage
period
and
then
on
the
day
of
sacrifice.
Body
weights
of
F0
female
rats
were
recorded
weekly
during
the
pre­
and
cohabitation
periods
and
then
on
gestation
days
(
GD)
0,
7,
10,
14,
18,
21,
and
25
(
if
necessary)
and
on
lactation
days
(
LD)
1,
5,
8,
11,
15,
and
22
(
terminal
body
weight).
Food
consumption
values
in
F0
male
rats
were
recorded
weekly
during
the
treatment
period,
while
in
F0
female
rats,
values
were
recorded
weekly
during
the
precohabitation
period,
on
GDs
0,
7,
10,
14,
18,
21,
and
25
and
on
LDs
1,
5,
8,
11,
and
15.
65
Estrous
cycling
was
evaluated
daily
by
examination
of
vaginal
cytology
beginning
21
days
before
the
scheduled
cohabitation
period
and
continuing
until
confirmation
of
mating
by
the
presence
of
sperm
in
a
vaginal
smear
or
confirmation
of
a
copulatory
plug.
On
the
day
of
scheduled
sacrifice,
the
stage
of
the
estrous
cycle
was
assessed.

For
mating,
one
male
rat
and
one
female
rat
per
group
were
cohabitated
for
a
maximum
of
14
days.
Female
rats
with
evidence
of
sperm
in
a
vaginal
smear
or
copulatory
plug
were
designated
as
GD
0
and
assigned
to
individual
housing.
Parental
females
were
evaluated
for
length
of
gestation,
fertility
index,
gestation
index,
number
and
sex
of
offspring
per
litter,
number
of
implantation
sites,
general
condition
of
the
dam
and
litter
during
the
postpartum
period,
litter
size
and
viability,
viability
index,
lactation
index,
percent
survival,
and
sex
ratio.
Maternal
behavior
of
the
dams
was
recorded
on
LDs
1,
5,
8,
15,
and
22.

F0
generation
animals
were
sacrificed
by
carbon
dioxide
asphyxiation
(
day
106
to
110
of
the
study
for
male
rats,
i.
e.,
after
completion
of
the
cohabitation
period;
and
LD
22
for
female
rats),
necropsied,
and
examined
for
gross
lesions.
Gross
necropsy
included
examination
of
external
surfaces
and
orifices,
as
well
as
internal
examination
of
tissues
and
organs.
Individual
organs
were
weighed
and
organ­
to­
body
weight
and
organ­
to­
brain
weight
ratios
were
calculated
for
the
brain,
kidneys,
spleen,
ovaries,
testes,
thymus,
liver,
adrenal
glands,
pituitary,
uterus
with
oviducts
and
cervix,
left
epididymis
(
whole
and
cauda),
right
epididymis,
prostate
and
seminal
vesicles,
(
with
coagulating
glands
and
with
and
without
fluid).
Tissues
retained
in
neutral
buffered
10%
formalin
for
possible
histological
evaluation
included
the
pituitary,
adrenal
glands,
vagina,
uterus,
with
oviducts,
cervix
and
ovaries,
right
testis,
seminal
vesicles,
right
epididymis,
and
prostate.
Histological
examination
was
performed
on
tissues
from
10
randomly
selected
rats
per
sex
from
the
control
and
high
dosage
groups.
All
gross
lesions
were
examined
histologically.
All
F0
generation
rats
that
died
or
appeared
moribund
were
also
examined.

Histological
examination
of
the
reproductive
organs
in
the
low­
and
mid­
dose
groups
was
conducted
in
rats
that
exhibited
reduced
fertility
by
either
failing
to
mate,
conceive,
sire,
or
deliver
healthy
offspring;
or
for
which
estrous
cyclicity
or
sperm
number,
motility,
or
morphology
were
altered.
Sperm
number,
motility,
and
morphology
were
evaluated
in
the
left
cauda
epididymis
of
F0
generation
male
rats;
testicular
spermatid
concentrations
were
evaluated
in
the
left
testis.
The
number
and
distribution
of
implantation
sites
were
recorded
in
F0
generation
female
rats.
Rats
that
did
not
deliver
a
litter
were
sacrificed
on
GD
25
and
examined
for
pregnancy
status.
Uteri
of
apparently
nonpregnant
rats
were
examined
to
confirm
the
absence
of
implantation
sites.
A
gross
necropsy
of
the
thoracic,
abdominal
and
pelvic
viscera
was
performed.
Female
rats
without
a
confirmed
mating
date
that
did
not
deliver
a
litter
were
sacrificed
on
an
estimated
day
25
of
gestation.

At
scheduled
sacrifice,
after
completion
of
the
cohabitation
period
in
F0
male
rats
and
on
LD
22
in
F0
female
rats,
blood
samples
(
10
males
and
10
females
each
for
the
10
and
30
mg/
kg/
day
dose
groups;
3
males
and
3
females
for
the
control
group)
were
collected
and
frozen
for
future
analysis.
The
methods
section
cites
that
liver
samples
were
also
collected,
but
no
other
details
were
provided
and
the
results
did
not
appear
to
be
available
at
the
time
of
the
report.
66
The
F1
generation
pups
in
each
litter
were
counted
once
daily.
Physical
signs
(
including
variations
from
expected
lactation
behavior
and
gross
external
physical
anomalies)
were
recorded
for
the
pups
each
day.
Pup
body
weights
were
recorded
on
LDs
1,
5,
8,
15
and
22.
On
LD
12,
all
F1
generation
male
pups
were
examined
for
the
presence
of
nipples.
Pups
that
died
before
examination
of
the
litter
for
pup
viability
on
LD
1
were
evaluated
for
vital
status
at
birth.
Pups
found
dead
on
LDs
2
to
22
were
examined
for
gross
lesions
and
for
the
cause
of
death.
All
F1
generation
rats
were
weaned
on
LD
22
based
on
observed
growth
and
viability
of
these
pups.

At
weaning
(
LD
22),
two
F1
generation
pups
per
sex
per
litter
per
group
(
60
male
and
60
female
pups
per
group)
were
selected
for
continued
evaluation,
resulting
in
600
total
rats
(
300
rats
per
sex)
assigned
to
the
five
dosage
groups.
At
least
two
male
pups
and
two
female
pups
per
litter,
when
possible,
were
selected.
F1
generation
pups
not
selected
for
continued
observation
for
sexual
maturation
were
sacrificed.
Three
pups
per
sex
per
litter
were
examined
for
gross
lesions.
Necropsy
included
a
single
cross­
section
of
the
head
at
the
level
of
the
frontal­
parietal
suture
and
examination
of
the
cross­
sectioned
brain
for
apparent
hydrocephaly.
The
brain,
spleen
and
thymus
from
one
of
the
three
selected
pups
per
sex
per
litter
were
weighed
and
the
brain,
spleen,
and
thymus
from
the
three
selected
pups
per
sex
per
litter
were
retained
for
possible
histological
evaluation.
All
remaining
pups
were
discarded
without
further
examination.

The
F1
generation
rats
were
given
the
same
dosage
level
of
the
test
substance
and
in
the
same
manner
as
their
respective
F0
generation
sires
and
dams.
Dosages
were
given
once
daily,
beginning
at
weaning
and
continuing
until
the
day
before
sacrifice.
F1
generation
female
rats
were
examined
for
age
of
vaginal
patency,
beginning
on
day
28
postpartum
(
LD
28).
F1
generation
male
rats
were
evaluated
for
age
of
preputial
separation,
beginning
on
day
39
postpartum
(
LD
39).
Body
weights
were
recorded
when
rats
reached
sexual
maturation.

Estrous
cycling
was
evaluated
daily
by
examination
of
vaginal
cytology
beginning
21
days
before
the
scheduled
cohabitation
period
and
continuing
until
confirmation
of
mating
by
the
presence
of
sperm
in
a
vaginal
smear
or
confirmation
of
a
copulatory
plug.
On
the
day
of
scheduled
sacrifice,
the
stage
of
the
estrous
cycle
was
assessed.

A
table
of
random
units
was
used
to
assign
F1
generation
rats
to
cohabitation,
one
male
rat
per
female
rat.
If
random
assignment
to
cohabitation
resulted
in
the
pairing
of
F1
generation
siblings,
an
alternate
assignment
was
made.
The
cohabitation
period
consisted
of
a
maximum
of
14
days.

Body
weights
of
the
F1
generation
male
rats
were
recorded
weekly
during
the
postweaning
period
and
on
the
day
of
sacrifice.
Body
weights
of
the
F1
generation
female
rats
were
recorded
weekly
during
the
postweaning
period
to
cohabitation,
and
on
DGs
0,
7,
10,
14,
18,
21
and
25
(
if
necessary)
and
on
LDs
1,
5,
8,
11,
15
and
22.
Food
consumption
values
for
the
F1
generation
male
rats
were
recorded
weekly
during
the
dosage
period.
Food
consumption
values
for
the
F1
generation
female
rats
were
recorded
weekly
during
the
postweaning
period
to
cohabitation,
on
GDs
0,
7,
10,
14,
18,
21
and
25
and
on
LDs
1,
5,
8,
11
and
15.
Because
pups
begin
to
consume
67
maternal
food
on
or
about
LD
15,
food
consumption
values
were
not
tabulated
after
LD
15.

At
scheduled
sacrifice,
the
F1
animals
were
subjected
to
gross
necropsy,
and
selected
organs
were
weighed
and
examined
histologically
as
described
above
for
the
F0
animals.
Sperm
analyses
were
also
conducted
as
described
for
the
F0
animals.

F2
generation
litters
were
examined
after
delivery
to
identify
the
number
and
sex
of
pups,
stillbirths,
live
births
and
gross
alterations.
Each
litter
was
evaluated
for
viability
at
least
twice
each
day
of
the
22­
day
postpartum
period.
Dead
pups
observed
at
these
times
were
removed
from
the
nesting
box.
Anogenital
distance
was
measured
for
all
live
F2
generation
pups
on
LDs
1
and
22.

Parental
Males
(
F0)

One
F0
male
rat
in
the
30
mg/
kg/
day
dose
group
was
sacrificed
on
day
45
of
the
study
due
to
adverse
clinical
signs
(
emaciation,
cold­
to­
touch,
and
decreased
motor
activity).
Necroscopic
examination
in
that
animal
revealed
a
pale
and
tan
liver,
and
red
testes.
All
other
F0
generation
male
rats
survived
to
scheduled
sacrifice.
Statistically
significant
increases
in
clinical
signs
were
also
observed
in
male
rats
in
the
high­
dose
group
that
included
dehydration,
urine­
stained
abdominal
fur,
and
ungroomed
coat.

Significant
reductions
in
body
weight
and
body
weight
gain
were
reported
for
most
of
the
dosage
period
and
continuing
until
termination
of
the
study
in
the
3,
10,
and
30
mg/
kg/
day
dose
groups.
Absolute
food
consumption
values
were
also
significantly
reduced
during
these
periods
at
the
30
mg/
kg/
day
dose
group,
while
significant
increases
in
relative
food
consumption
values
were
observed
in
the
3,
10,
and
30
mg/
kg/
day
within
those
same
periods.

No
treatment­
related
effects
were
reported
at
any
dose
level
for
any
of
the
mating
and
fertility
parameters
assessed,
including
numbers
of
days
to
inseminate,
numbers
of
rats
that
mated,
fertility
index,
numbers
of
rats
with
confirmed
mating
dates
during
the
first
and
second
week
of
cohabitation,
and
numbers
of
pregnant
rats
per
rats
in
cohabitation.
At
necropsy,
none
of
the
sperm
parameters
evaluated
(
sperm
number,
motility,
or
morphology)
were
affected
by
treatment
at
any
dose
level.

At
necropsy,
statistically
significant
reductions
in
terminal
body
weights
were
seen
at
3,
10,
and
30
mg/
kg/
day.
Absolute
weights
of
the
left
and
right
epididymides,
left
cauda
epididymis,
seminal
vesicles
(
with
and
without
fluid),
prostate,
pituitary,
left
and
right
adrenals,
spleen,
and
thymus
were
also
significantly
reduced
at
30
mg/
kg/
day.
The
absolute
weight
of
the
seminal
vesicles
without
fluid
was
significantly
reduced
in
the
10
mg/
kg/
day
dose
group.
The
absolute
weight
of
the
liver
was
significantly
increased
in
all
dose­
groups.
Kidney
weights
were
significantly
increased
in
the
1,
3,
and
10
mg/
kg/
day
dose
groups,
but
significantly
decreased
in
the
30
mg/
kg/
day
group.
All
organ
weight­
to­
terminal
body
weight
and
ratios
were
significantly
increased
in
all
treated
groups.
Organ
weight­
to­
brain
weight
ratios
were
significantly
reduced
for
some
organs
at
the
high
dose
group,
and
significantly
increased
for
other
organs
among
all
68
treated
groups.

No
treatment­
related
effects
were
seen
at
necropsy
or
upon
microscopic
examination
of
the
reproductive
organs,
with
the
exception
of
increased
thickness
and
prominence
of
the
zona
glomerulosa
and
vacuolation
of
the
cells
of
the
adrenal
cortex
in
the
10
and
30
mg/
kg/
day
dose
groups.

Serum
analysis
for
the
F0
generation
males
sampled
at
the
end
of
cohabitation
showed
that
PFOA
was
present
in
all
samples
tested,
including
controls.
Control
males
had
an
average
concentration
of
0.0344+
0.0148
ug/
ml
PFOA.
Treated
males
had
51.1+
9.30
and
45.3+
12.6
ug/
ml,
respectively
for
the
10
and
30
mg/
kg/
day
dose
groups.

Parental
Females
(
F0)

No
treatment­
related
deaths
or
adverse
clinical
signs
were
reported
in
parental
females
at
any
dose
level.
No
treatment­
related
effects
were
reported
for
body
weights,
body
weight
gains,
and
absolute
and
relative
food
consumption
values.

There
were
no
treatment­
related
effects
on
estrous
cyclicity,
mating
or
fertility
parameters.
None
of
the
natural
delivery
and
litter
observations
were
affected
by
treatment,
that
is,
the
numbers
of
dams
delivering
litters,
the
duration
of
gestation,
the
averages
for
implantation
sites
per
delivered
litter,
the
gestation
index
(
number
of
dams
with
one
or
more
liveborn
pups/
number
of
pregnant
rats),
the
numbers
of
dams
with
stillborn
pups,
dams
with
all
pups
dying,
liveborn
and
stillborn
pups
viability
index,
pup
sex
ratios,
and
mean
birth
weights
were
comparable
to
controls
among
all
treated
groups.

Necropsy
and
histopathological
evaluation
were
also
unremarkable.
Terminal
body
weights,
organ
weights,
and
organ­
to­
terminal
body
weight
ratios
were
comparable
to
control
values
for
all
treated
groups,
except
for
kidney
and
liver
weights.
The
weights
of
the
left
and
right
kidney,
and
the
ratios
of
these
organ
weights­
to­
terminal
body
weights,
and
of
the
left
kidney
weight­
tobrain
weight
were
significantly
reduced
at
the
highest
dose
of
30
mg/
kg/
day.
The
ratio
of
liver
weights­
to­
terminal
body
weight
was
also
significantly
reduced
at
3
and
10
mg/
kg/
day.

Results
of
the
serum
analysis
in
F0
generation
females
sampled
on
LD
22
showed
that
PFOA
was
present
in
all
samples
tested,
except
in
controls
where
the
level
was
below
the
limits
of
quantitation
(
0.00528ugm/
l).
Treated
females
had
an
average
concentration
of
0.37+
0.0805
and
1.02+
0.425
ug/
ml,
respectively
for
the
10
and
30
mg/
kg/
day
dose
groups.

F1
Generation
­
Males
No
effects
were
reported
at
any
dose
level
for
the
viability
and
lactation
indices.
No
differences
between
treated
and
control
groups
were
noted
for
the
numbers
of
pups
surviving
per
litter,
the
percentage
of
male
pups,
litter
size
and
average
pup
body
weight
per
litter
at
birth.
Pup
body
weight
on
a
per
litter
basis
(
sexes
combined)
was
significantly
reduced
in
the
30
mg/
kg/
day
69
group
on
days
1,
5,
and
8
of
lactation.
Of
the
pups
necropsied
at
weaning,
no
statistically
significant,
treatment­
related
differences
were
observed
for
the
weights
of
the
brain,
spleen
and
thymus
and
the
ratios
of
these
organ
weights
to
the
terminal
body
weight
and
brain
weight.

Significant
increases
in
treatment­
related
deaths
(
7
animals
total)
were
reported
in
F1
males
in
the
high
dose
group
of
30
mg/
kg/
day.
One
rat
was
moribund
sacrificed
on
day
39
postweaning
and
another
was
found
dead
on
day
107
postweaning,
but
the
majority
of
the
F1
male
rats
were
found
dead
on
days
2­
4
postweaning.

Statistically
significant
increases
in
clinical
signs
of
toxicity
were
also
observed
in
F1
males
during
most
of
entire
postweaning
period.
These
signs
included
an
increased
incidence
of
annular
constriction
of
the
tail
at
all
doses,
with
statistical
significance
at
the
1,
10,
and
30
mg/
kg/
day;
a
significant
increase
at
10
and
30
mg/
kg/
day
in
the
number
of
male
rats
that
were
emaciated;
and
a
significant
increase
in
the
incidence
of
urine­
stained
abdominal
fur,
decreased
motor
activity,
and
abdominal
distention
at
30
mg/
kg/
day.

Body
weights
and
body
weight
gains
were
statistically
significantly
reduced
prior
to
and
during
cohabitation
and
during
the
entire
dosing
period
in
all
treated
groups.
Statistically
significant
reductions
in
body
weights
were
observed
at
10
and
30
mg/
kg/
day
during
days
8­
15,
22­
29,
29­
36,
43­
50,
and
50­
57
postweaning.
Body
weight
gains
were
also
significantly
reduced
in
the
30
mg/
kg/
day
group
on
days
1­
8,
15­
22,
36­
43,
57­
64,
and
64­
70
postweaning.
Statistically
significant,
dose­
related
reductions
in
body
weight
gains
were
observed
for
the
entire
dosage
period
(
days
1­
113
postweaning).
Absolute
food
consumption
values
were
significantly
reduced
at
10
and
30
mg/
kg/
day
during
the
entire
precohabitation
period
(
days
1­
70
postweaning),
while
relative
food
consumption
values
were
significantly
increased.

Statistically
significant
(
p<
0.01)
delays
in
sexual
maturation
(
the
average
day
of
preputial
separation)
were
observed
in
high­
dose
animals
versus
concurrent
controls
(
52.2
days
of
age
versus
48.5
days
of
age,
respectively).

No
apparent
effects
were
observed
on
any
of
the
mating
or
fertility
parameters
including
fertility
and
pregnancy
indices
(
number
of
pregnancies
per
number
of
rats
that
mated
and
rats
in
cohabitation,
respectively),
the
number
of
days
to
inseminate,
the
number
of
rats
that
mated,
and
the
number
of
rats
with
confirmed
mating
dates
during
the
first
week.
No
statistically
significant,
treatment­
related
effects
were
observed
on
any
of
the
sperm
parameters
(
motility,
concentration,
or
morphology).

Necroscopic
examination
revealed
statistically
significant
treatment­
related
effects
at
3,
10,
and
30
mg/
kg/
day
ranging
from
tan
areas
in
the
lateral
and
median
lobes
of
the
liver
to
moderate
to
slight
dilation
of
the
pelvis
of
one
or
both
kidneys.

Statistically
significant,
dose­
related
decreases
in
terminal
body
weights
of
parental
F1
males
were
observed.
The
absolute
weights
of
the
liver
were
significantly
increased
and
the
absolute
weights
of
the
spleen
were
significantly
decreased
at
all
treated
groups.
The
absolute
weights
of
70
the
left
and/
or
right
kidneys
were
significantly
increased
in
the
1
and
3
mg/
kg/
day
dose
groups
and
significantly
decreased
in
the
30
mg/
kg/
day
dose
group.
The
absolute
weight
of
the
thymus
was
also
significantly
decreased
in
the
10
and
30
mg/
kg/
day
dose
groups.
The
absolute
weight
of
the
prostate,
brain
and
left
adrenal
gland
were
significantly
decreased
in
the
30
mg/
kg/
day
dosage
group.
The
ratios
of
the
weights
of
the
seminal
vesicles,
with
and
without
fluid,
liver
and
left
and
right
kidneys
to
the
terminal
body
weights
were
significantly
increased
in
all
treated
groups.
The
ratios
of
the
weights
of
the
left
testis,
with
and
without
the
tunica
albuginea
and
the
right
testis
to
the
terminal
body
weight,
were
significantly
increased
at
3
mg/
kg/
day
and
higher.
The
ratios
of
the
weights
of
the
left
epididymis,
left
cauda
epididymis,
right
epididymis
and
brain
to
the
terminal
body
weight
were
significantly
increased
at
10
mg/
kg/
day
and
higher.
The
ratios
of
the
weight
of
the
seminal
vesicles
with
fluid
to
the
brain
weight
were
increased
at
1
mg/
kg/
day
and
higher,
with
statistical
significance
at
1
and
10
mg/
kg/
day.
The
ratios
of
the
liver
weight­
to­
brain
weight
were
significantly
increased
in
the
1
mg/
kg/
day
and
higher
dosage
groups,
and
the
ratios
of
the
left
and
right
kidney
weights­
to­
brain
weight
were
significantly
increased
in
all
treated
groups.
The
ratios
of
the
spleen
weight­
to­
brain
weight
were
significantly
decreased
at
1
mg/
kg/
day
and
higher,
and
the
ratios
of
the
thymus
weight­
to­
brain
weight
were
significantly
decreased
at
10
and
30
mg/
kg/
day.
The
ratios
of
the
left
and
right
testes
weight­
tobrain
weight
were
increased
in
the
3
mg/
kg/
day
and
higher
dosage
groups.
These
ratios
were
significantly
increased
at
10
mg/
kg/
day
(
right
testis
only)
and
30
mg/
kg/
day.

Histopathologic
examination
of
the
reproductive
organs
was
unremarkable;
however,
treatmentrelated
microscopic
changes
were
observed
in
the
adrenal
glands
of
high­
dose
animals
(
cytoplasmic
hypertrophy
and
vacuolation
of
the
cells
of
the
adrenal
cortex)
and
in
the
liver
of
animals
treated
with
3,
10,
and
30
mg/
kg/
day
(
hepatocellular
hypertrophy).
No
other
treatmentrelated
effects
were
reported.

F1
Generation
 
Females
No
effects
were
reported
at
any
dose
level
for
the
viability
and
lactation
indices.
No
differences
between
treated
and
control
groups
were
noted
for
the
numbers
of
pups
surviving
per
litter,
the
percentage
male
pups,
litter
size
and
average
pup
body
weight
per
litter
at
birth.
Pup
body
weight
on
a
per
litter
basis
(
sexes
combined)
was
significantly
reduced
in
the
30
mg/
kg/
day
group
on
days
1,
5,
and
8
of
lactation.

At
30
mg/
kg/
day,
one
pup
from
one
dam
died
prior
to
weaning
on
lactation
day
1
(
LD1).
Additionally,
on
lactation
days
6
and
8,
statistically
significant
increases
in
the
numbers
of
pups
found
dead
were
observed
at
3
and
30
mg/
kg/
day.
According
to
the
study
authors,
this
was
not
considered
to
be
treatment
related
because
they
did
not
occur
in
a
dose­
related
manner
and
did
not
appear
to
affect
any
other
measures
of
pup
viability
that
included,
numbers
of
surviving
pups
per
litter
and
live
litter
size
at
weighing.
An
independent
statistical
analysis
was
conducted
by
US
EPA
(
2002b).
No
significant
differences
were
noted
between
dose
groups
and
there
was
no
significant
trend.
71
Of
the
pups
necropsied
at
weaning,
no
statistically
significant,
treatment­
related
differences
were
observed
for
the
weights
of
the
brain,
spleen
and
thymus
and
the
ratios
of
these
organ
weights
to
the
terminal
body
weight
and
brain
weight.

An
increase
in
treatment­
related
mortality
(
6
animals
total)
was
observed
in
F1
females
on
postweaning
days
2­
8
at
the
highest
dose
of
30
mg/
kg/
day.
No
adverse
clinical
signs
of
treatment­
related
toxicity
were
reported
for
any
dose
level
during
any
time
of
the
study
period.

Statistically
significant
decreases
in
body
weights
and
body
weight
gains
were
observed
in
highdose
animals
on
days
8,
15,
22,
29,
50,
and
57
postweaning,
during
precohabitation
(
recorded
on
the
day
cohabitation
began,
when
F1
generation
rats
were
92­
106
days
of
age),
and
during
gestation
and
lactation.
Decreases
in
absolute
food
consumption
were
observed
during
days
1­
8,
8­
15
postweaning
during
precohabitation
and
during
gestation
and
lactation
in
animals
treated
with
30
mg/
kg/
day.
Relative
food
consumption
values
were
comparable
across
all
treated
groups.

Statistically
significant
(
p<
0.01)
delays
in
sexual
maturation
(
the
average
day
of
vaginal
patency)
were
observed
in
high­
dose
animals
versus
concurrent
controls
(
36.6
days
of
age
versus
34.9
days
of
age,
respectively).

Prior
to
mating,
the
study
authors
noted
a
statistically
significant
increase
in
the
average
numbers
of
estrous
stages
per
21
days
in
high­
dose
animals
(
5.4
versus
4.7
in
controls).
For
this
calculation,
the
number
of
independent
occurrences
of
estrus
in
the
21
days
of
observation
was
determined.
This
type
of
calculation
can
be
used
as
a
screen
for
effects
on
the
estrous
cycle,
but
a
more
detailed
analysis
should
then
be
conducted
to
determine
whether
there
is
truly
an
effect.
3M
Company
(
2002)
recently
completed
an
analysis
that
showed
there
were
no
effects
on
the
estrous
cycle;
there
were
no
differences
in
the
number
of
females
with
>
3
days
of
estrus
or
with
>
4
days
of
diestrus
in
the
control
and
high
dose
groups.
Analyses
conducted
by
the
US
EPA
(
2002a)
also
demonstrated
that
there
were
no
differences
in
the
estrous
cycle
among
the
control
and
high
dose
groups.
The
cycles
were
evaluated
as
having
either
regular
4­
5
day
cycles,
uneven
cycling
(
defined
as
brief
periods
with
irregular
pattern)
or
periods
of
prolonged
diestrus
(
defined
as
4­
6
day
diestrus
periods)
extended
estrus
(
defined
as
3
or
4
days
of
cornified
smears),
possibly
pseudopregnant,
(
defined
as
6­
greater
days
of
leukocytes)
or
persistent
estrus
(
defined
as
5­
or
greater
days
of
cornified
smears).
The
two
groups
were
not
different
in
any
of
the
parameters
measured.
Thus,
the
increase
in
the
number
of
estrous
stages
per
21
days
that
was
noted
by
the
study
authors
is
due
to
the
way
in
which
the
calculation
was
done,
and
is
not
biologically
meaningful.

No
effects
on
any
of
the
mating
and
fertility
parameters
(
numbers
of
days
in
cohabitation,
numbers
of
rats
that
mated,
fertility
index,
rats
with
confirmed
mating
dates
during
the
first
week
of
cohabitation
and
number
of
rats
pregnant
per
rats
in
cohabitation).

All
natural
delivery
observations
were
unaffected
by
treatment
at
any
dose
level.
Numbers
of
dams
delivering
litters,
the
duration
of
gestation,
averages
for
implantation
sites
per
delivered
72
litter,
the
gestation
index
(
number
of
dams
with
one
or
more
liveborn
pups/
number
of
pregnant
rats),
the
numbers
of
dams
with
stillborn
pups,
dams
with
all
pups
dying
and
liveborn
and
stillborn
pups
were
comparable
among
treated
and
control
groups.

No
treatment­
related
effects
were
observed
on
terminal
body
weights.
The
absolute
weight
of
the
pituitary
and
the
ratios
of
the
pituitary
weight­
to­
terminal
body
weight
and
to
the
brain
weight
were
significantly
decreased
at
3
mg/
kg/
day
and
higher,
but
did
not
show
a
doseresponse
No
other
differences
were
reported
for
the
absolute
weights
or
ratios
for
other
organs
evaluated.
No
treatment­
related
effects
were
reported
following
necroscopic
and
histopathologic
examinations.

F2
Generation
Offspring
No
treatment­
related
adverse
clinical
signs
were
observed
at
any
dose
level.
Likewise,
no
treatment­
related
effects
were
reported
following
necroscopic
examination,
with
the
exception
of
no
milk
in
stomach
in
pups
that
were
found
dead.
The
numbers
of
pups
found
either
dead
or
stillborn
did
not
show
a
dose­
response
(
3/
28,
6/
28,
10/
28,
10/
28,
and
6/
28
in
0,
1,
3,
10,
and
30
mg/
kg/
day
dose
groups,
respectively)
and
therefore
were
unlikely
related
to
treatment.

No
effects
were
reported
at
any
dose
level
for
the
viability
and
lactation
indices.
No
differences
between
treated
and
control
groups
were
noted
for
the
numbers
of
pups
surviving
per
litter,
the
percentage
of
male
pups,
litter
size
and
average
pup
body
weight
per
litter
when
measured
on
LDs
1,
5,
8,
15,
or
22.
Anogenital
distances
measured
for
F2
male
and
female
pups
on
LDs
1
and
22
were
also
comparable
among
the
five
dosage
groups
and
did
not
differ
significantly.

Statistically
significant
increases
(
p<
0.01)
in
the
number
of
pups
found
dead
were
observed
on
lactation
day
1
at
the
3
and
10
mg/
kg/
day
dosage
groups.
According
to
the
study
authors,
this
was
not
considered
to
be
treatment
related
because
they
did
not
occur
in
a
dose­
related
manner
and
did
not
appear
to
affect
any
other
measures
of
pup
viability
that
included,
numbers
of
surviving
pups
per
litter
and
live
litter
size
at
weighing.
An
independent
statistical
analysis
was
conducted
by
US
EPA
(
2002b).
No
significant
differences
were
observed
between
dose
groups
and
there
was
no
significant
trend.

Terminal
body
weights
in
F2
pups
were
not
significantly
different
from
controls.
Absolute
weights
of
the
brain,
spleen
and
thymus
and
the
ratios
of
these
organ
weights­
to­
terminal
body
weight
and
to
brain
weight
were
also
comparable
among
treated
and
control
groups.

Conclusions
Dosing
with
APFO
at
30
mg/
kg/
day
appeared
to
delay
the
onset
of
sexual
maturation
in
both
male
and
female
F1
offspring.
The
authors
of
the
study
contend
that
the
delays
in
sexual
maturation
(
preputial
separation
or
vaginal
patency)
observed
in
high­
dose
animals
are
due
to
the
fact
that
these
animals
have
a
decreased
gestational
age,
a
variable
which
they
have
defined
as
the
time
in
days
from
evidence
of
mating
in
the
F0
generation
until
evidence
of
sexual
73
maturation
in
the
F1
generation.
The
authors
state
that
gestational
age
appeared
to
be
decreased
in
high­
dose
animals
at
the
time
of
acquisition
(
the
time
when
sexual
maturation
was
reached),
which
they
believe
meant
the
animals
in
that
group
were
younger
and
more
immature
than
the
control
group,
in
which
there
was
no
significant
difference
in
sexual
maturation.

In
order
to
test
this
hypothesis,
the
authors
covaried
separately
the
decreases
in
body
weight
and
in
gestational
age
with
the
delays
in
sexual
maturation
in
order
to
determine
whether
or
not
body
weights
and
gestational
age
were
a
contributing
factor.
When
the
body
weight
was
covaried
with
the
time
to
sexual
maturation,
the
time
to
sexual
maturation
showed
a
dose
related
delay
that
was
statistically
significant
at
the
p<
0.05.
This
suggests
that
the
delay
in
sexual
maturation
was
partly
related
to
body
weight,
but
not
entirely.
When
gestational
age
was
covaried
with
the
time
to
sexual
maturation,
there
was
no
significant
difference
in
the
time
of
onset
of
sexual
maturation
between
controls
and
high­
dose
animals.
This
indicates
that
the
effect
of
delayed
sexual
maturation
could
possibly
be
attributed
to
decreased
gestational
age.

While
it
is
known
and
commonly
accepted
that
changes
in
the
body
weights
of
offspring
can
affect
the
time
to
sexual
maturation,
whether
or
not
gestational
age,
as
defined
by
the
authors,
also
affects
the
time
of
acquisition
is
purely
speculative,
especially
since
there
was
no
data
provided
by
the
authors
to
support
this
relationship.
Additionally,
covarying
gestational
age
with
time
to
sexual
maturation
is
problematic
from
a
statistical
standpoint.
Since
there
was
no
significant
change
in
the
length
of
gestation
at
30
mg/
kg/
day,
based
on
the
authors'
definition
of
`
gestational
age',
the
decreases
in
gestational
age
would
have
to
be
due
mostly
to
changes
in
time
to
sexual
maturation.
Therefore,
sexual
maturation
is
essentially
being
covaried
with
itself.
Still,
even
if
a
relationship
between
gestational
age
and
time
to
sexual
maturation
were
shown,
it
merely
offers
an
explanation
for
the
observed
delays
in
sexual
maturation
in
high­
dose
animals,
but
does
not
diminish
its
significance.

Therefore,
under
the
conditions
of
the
study,
the
LOAEL
for
F0
parental
males
is
considered
to
be
1
mg/
kg/
day,
the
lowest
dose
tested,
based
on
significant
increases
in
the
liver
and
kidney
weights­
to­
terminal
body
weight
and
to
brain
weight
ratios.
A
NOAEL
for
the
F0
parental
males
could
not
be
determined
since
treatment­
related
effects
were
seen
at
all
doses
tested.

The
NOAEL
and
LOAEL
for
F0
parental
females
are
considered
to
be
10
and
30
mg/
kg/
day,
respectively,
based
on
significant
reductions
in
kidney
weight
and
kidney
weight­
to­
terminal
body
weight
and
to
brain
weight
ratios
observed
at
the
highest
dose.

The
LOAEL
for
F1
generation
males
is
considered
to
be
1
mg/
kg/
day,
based
on
significant
decreases
in
body
weights
and
body
weight
gains,
and
in
terminal
body
weights;
and
significant
changes
in
absolute
liver
and
spleen
weights
and
in
the
ratios
of
liver,
kidney,
and
spleen
weights­
to­
brain
weights;
and
based
on
significant,
dose­
related
reductions
in
body
weights
and
body
weight
gains
observed
prior
to
and
during
cohabitation
and
during
the
entire
dosing
period.
A
NOAEL
for
the
F1
males
could
not
be
determined
since
treatment­
related
effects
were
seen
at
all
doses
tested.
74
The
NOAEL
and
LOAEL
for
F1
generation
females
are
considered
to
be
10
and
30
mg/
kg/
day,
respectively,
based
on
statistically
significant
increases
in
postweaning
mortality,
delays
in
sexual
maturation
(
time
to
vaginal
patency),
decreases
in
body
weight
and
body
weight
gains,
and
decreases
in
absolute
food
consumption,
all
observed
at
the
highest
dose
tested.

The
NOAEL
for
the
F2
generation
offspring
was
considered
to
be
30
mg/
kg/
day.
No
treatmentrelated
effects
were
observed
at
any
doses
tested
in
the
study.
However,
it
should
be
noted
that
the
F2
pups
were
sacrificed
at
weaning,
and
thus
it
was
not
possible
to
ascertain
the
potential
post­
weaning
effects
that
were
noted
in
the
F1
generation.

3.8
Carcinogenicity
Studies
in
Animals
3.8.1
Cancer
Bioassays
The
carcinogenic
potential
of
APFO
has
been
investigated
in
a
two­
year
feeding
study
in
rats
(
3M,
1987).
In
this
study,
groups
of
50
male
and
50
female
Sprague­
Dawley
(
Crl:
CD
BR)
rats
were
fed
diets
containing
0,
30
or
300
ppm
APFO
for
two
years.
Groups
of
15
additional
rats
per
sex
were
fed
0,
or
300
ppm
APFO
and
evaluated
at
the
one­
year
interim
sacrifice.
The
mean
actual
test
article
consumption
was:
males,
1.3
and
14.2
mg/
kg/
day;
females,
1.6
and
16.1
mg/
kg/
day
for
the
low
and
high­
dose
groups,
respectively.

There
was
a
dose­
related
decrease
in
body
weight
gain
in
the
male
rats
and
to
a
lesser
extent,
in
the
female
rats
as
compared
to
the
controls;
the
decreases
were
statistically
significant
in
the
high­
dose
groups
of
both
sexes.
The
body
weight
changes
are
treatment
related
since
feed
consumption
was
actually
increased
(
rather
than
decreased).
There
were
no
differences
in
mortality
between
the
treated
and
untreated
groups;
the
survival
rates
at
the
end
of
104
weeks
for
the
control,
low­,
and
high­
dose
groups
were:
male,
70%,
72%
and
88%;
females,
50%,
48%
and
58%.
The
only
clinical
sign
observed
was
a
dose­
related
increase
in
ataxia
in
the
female
rats;
the
incidences
in
the
control,
low­
and
high­
dose
groups
were:
4%,
18%
and
30%.
Significant
decreases
in
red
blood
cell
counts,
hemoglobin
concentrations
and
hematocrit
values
were
observed
in
the
high­
dose
male
and
female
rats
as
compared
to
control
values.
Clinical
chemistry
changes
indicative
of
liver
toxicity
included
increases
in
alanine
aminotransferase
(
ALT),
aspartate
aminotransferase
(
AST)
and
alkaline
phosphatase
(
AP)
in
both
treated
male
groups
from
3­
18
months,
but
only
in
the
high­
dose
males
at
24
months.
Increases
in
relative
liver
and
kidney
weights
were
noted
in
both
high­
dose
male
and
female
rats.
Significant
nonneoplastic
lesions
were
seen
primarily
in
the
liver
and
testis;
there
were
increases
in
the
incidence
of
liver
masses,
hyperplastic
nodules
and
foci,
and
in
testicular
masses
in
the
highdose
male
group.
Other
liver
toxic
effects
include
dose­
related
increases
in
the
incidence
of
diffuse
hepatomegalocytosis,
cystoid
degeneration,
and
portal
mononuclear
cell
infiltration
in
both
male
and
female
treated
groups;
these
increases
were
statistically
significant
in
the
highdose
males.
A
statistically
significant,
dose­
related
increase
in
the
incidence
of
ovarian
tubular
hyperplasia
was
found
in
female
rats;
the
incidence
of
this
lesion
in
the
control,
low­,
and
highdose
groups
was
0%,
14%,
and
32%,
respectively.
Based
on
these
toxic
effects,
the
high
dose
selected
in
this
study
appears
to
have
reached
the
Maximum
Tolerated
Dose
(
MTD).
Based
on
75
decreased
body
weight
gain,
increased
liver
and
kidney
weights
and
toxicity
in
the
hematological
and
hepatic
systems,
the
LOAEL
for
male
and
female
rats
is
300
ppm.
[
Based
on
increases
in
the
incidence
of
ataxia
and
ovarian
tubular
hyperplasia,
the
LOAEL
for
female
rats
is
30
ppm.]

At
the
termination
of
the
study,
a
slight
increase
in
the
incidence
of
various
neoplasms
(
tumors
of
the
liver,
testis,
thyroid,
adrenal
and
mammary
glands,
etc.)
was
seen
in
the
treated
animals.
Among
them,
the
increased
incidences
of
testicular
(
Leydig)
cell
adenomas
in
the
high­
dose
male
rats,
and
of
mammary
fibroadenoma
in
both
groups
of
female
rats
were
statistically
significant
(
P<
0.05)
as
compared
to
the
concurrent
controls.
The
incidence
of
the
Leydig
cell
tumors
(
LCT)
in
the
control,
low­
and
high­
dose
males
was
0/
50
(
0%),
2/
50
(
4%)
and
7/
50
(
14%),
respectively;
the
respective
incidences
of
mammary
fibroadenoma
in
the
female
groups
were
11/
50
(
22%),
21/
50
(
42%)
and
24/
50
(
48%).
The
increases
are
also
statistically
significant
as
compared
to
the
historical
control
incidences
(
LCT,
0.82%;
mammary
fibroadenoma,
19.0%)
observed
in
1,340
male
and
1,329
female
Sprague­
Dawley
control
rats
used
in
17
carcinogenicity
studies
(
Chandra
et
al.,
1992).
The
spontaneous
incidence
of
LCT
in
2­
year
old
Sprague­
Dawley
rats
in
other
studies
was
reported
to
be
approximately
5%
(
cited
in:
Clegg
et
al.,
1997).
Therefore,
under
the
conditions
of
this
study,
APFO
is
carcinogenic
in
Sprague­
Dawley
rats,
inducing
Leydig
cell
tumors
in
the
male
rats
and
mammary
fibroadenomas
in
the
female
rats.

The
induction
of
Leydig
cell
tumors
was
confirmed
in
a
follow­
up
2­
year
mechanism
study
of
PFOA
in
male
Sprague­
Dawley
(
CD)
rats
at
a
dietary
level
of
300
ppm
(
Cook
et
al.,
1994;
Biegel
et
al.
2001).
A
significantly
increased
LCT
incidence
was
observed
in
the
treated
rats
(
8/
76,
11%)
as
compared
to
the
controls
(
0/
80,
0%).
In
addition,
PFOA
also
caused
significantly
increased
incidences
of
liver
tumors
and
pancreatic
acinar
cell
tumors.
The
incidences
of
liver
adenomas
in
the
control
and
treated
groups
were
2/
80
(
3%)
and
10/
76
(
13%),
respectively,
whereas
those
for
the
pancreatic
acinar
cell
adenomas
were
0/
80
(
0%)
and
7/
76
(
9%).
There
was
one
pancreatic
acinar
cell
caricinoma
in
76
of
the
treated
rats
and
none
in
80
controls.
The
incidence
of
combined
pancreatic
acinar
cell
adenoma/
carcinoma
in
the
treated
rats
(
8/
76,
11%)
was
also
significantly
increased
as
compared
to
the
controls
(
0/
80,
0%).

PFOA
has
also
been
shown
to
promote
liver
carcinogenesis
in
rodents
(
Abdellatif
et
al.,
1991;
Nilsson
et
al.,
1991).

3.8.2
Mode
of
Action
Studies
The
mechanism(
s)
of
toxicological/
carcinogenic
action
of
PFOA
is
not
clearly
understood.
PFOA
was
not
mutagenic
in
the
Ames
test
using
five
strains
of
Salmonella
typhimurium,
or
in
an
assay
with
Saccharomyces
cerevisiae
(
Griffith
and
Long,
1980).
Short­
term
genotoxicity
assays
appear
to
suggest
that
PFOA
is
not
a
DNA­
reactive
compound.
However,
when
tested
with
metabolic
activation,
PFOA
induced
significant
increases
in
chromosomal
aberrations
and
in
polyploidy
in
CHO
cells
(
Murli,
1996).
The
significance
of
these
genotoxic
effects
is
unclear.
Available
data
appear
to
indicate
that
the
induction
of
tumors
by
PFOA
is
due
to
a
non­
genotoxic
mechanism,
involving
activation
of
receptors
and
perturbations
of
the
endocrine
system.
The
76
Agency
is
currently
examining
these
postulated
modes
of
action
in
detail.
The
following
summaries
are
not
meant
to
be
a
detailed
review
of
the
literature,
but
simply
summarize
the
current
scientific
evidence.

3.8.2.1
Liver
Tumors
It
has
been
well
documented
that
APFO
is
a
potent
peroxisome
proliferator,
inducing
peroxisome
proliferation
in
the
liver
of
rats
and
mice
(
e.
g.,
Ikeda
et
al.,
1985;
Pastoor
et
al.,
1987;
Sohlenius
et
al.,
1992).
A
sex­
related
difference
in
the
induction
of
liver
peroxisome
proliferation
exists
in
rats
(
Kawashima
et
al.,
1989),
but
not
in
mice
(
Sohlenius
et
al.,
1992).
The
higher
induction
of
liver
peroxisome
proliferation
in
male
rats
was
shown
to
be
strongly
dependent
on
the
sex
hormone
testosterone
(
Kawashima
et
al.,
1989).
Like
many
other
peroxisome
proliferators,
APFO
has
also
been
shown
to
cause
hepatomegaly
(
an
early
biomarker
of
peroxisome
proliferator
hepatocarcinogenesis)
in
rats
(
Takagi,
et
al.,
1992;
Cook,
1994)
and
mice
(
Kennedy,
1987),
and
induce
oxidative
DNA
damage
in
liver
of
rats
(
Takagi
et
al.,
1991).
The
totality
of
these
data
appears
to
suggest
that
the
liver
toxicity
and
carcinogenicity
of
APFO
may
be
related
to
induction
of
peroxisome
proliferation.
Meanwhile,
estrogen
has
been
shown
to
promote
hepatocarcinogenesis
in
rats
(
Yager
and
Yager,
1980;
Cameron
et
al.,
1982);
an
increase
in
estrogen
levels
after
APFO
exposure
(
discussed
below)
may
also
play
a
role
in
hepatocarcinogenesis
in
rats.
Recently,
IARC
(
1995)
concluded
that
the
liver
tumors
induced
in
rodents
by
PPAR­
alpha
agonists
are
unlikely
to
be
operative
in
humans
based
on
our
current
understanding
of
the
animal
mode
of
action.

3.8.2.2
Leydig
Cell
Tumors
A
large
number
of
non­
genotoxic
compounds
of
diverse
chemical
structures
have
been
reported
to
induce
Leydig
cell
tumors
(
LCT)
in
rats,
mice,
or
dogs.
A
review
of
the
available
information
on
LCT
induction
in
animals
led
a
workshop
panel
to
classify
these
compounds
into
seven
groups
based
on
their
modes
of
action
(
Clegg
et
al.,
1997).
The
common
theme
in
the
mode
of
action
for
most
compounds
is
that
these
compounds
affect
the
hormonal
control
of
Leydig
cell
growth
by
disrupting
the
hypothalamic­
pituitary­
testicular
axis
at
various
points
that
result
in
increasing
the
serum
levels
of
luteinizing
hormone
(
LH).
It
has
been
postulated
that
in
addition
to
stimulating
the
production
of
testosterone,
LH
may
also
play
a
mitogenic
role
in
the
Leydig
cells;
a
sustained
increase
in
circulating
LH
levels
and
chronic
stimulation
of
Leydig
cells
by
growth­
stimulating
mediators
such
as
IGF­
1,
TGF­
 ,
leukotrienes
and
various
free
radicals
can
lead
to
LCT
development
(
rev.
in:
Clegg
et
al.,
1997).

A
series
of
studies
have
been
conducted
to
investigate
the
mechanism
of
tumor
formation
in
male
Sprague­
Dawley
(
CD)
rats
exposed
to
APFO
(
Cook
et
al.,
1992;
Biegel
et
al.,
1995;
Liu
et
al.,
1996).
No
significant
increases
in
LH
were
seen
in
the
rats
after
treatment
of
APFO
at
various
dose
levels
for
14
days.
However,
serum
and
testicular
levels
of
estradiol
were
significantly
increased
and
testosterone
levels
were
significantly
decreased.
It
was
postulated
that
the
elevated
estradiol
levels
may
cause
Leydig
cell
hyperplasia
and
tumor
formation
by
acting
as
a
mitogen
and/
or
enhancing
growth
factor
secretion;
the
transforming
growth
factor
 
77
(
TGF
 ),
which
binds
to
the
epidermal
growth
factor
(
EGF)
receptor
and
stimulated
cell
proliferation,
for
instance,
has
been
detected
in
Leydig
cells
(
Teerds
et
al.,
1990).
Subsequent
experiments
have
shown
that
APFO
increased
the
levels
of
estradiol
by
inducing
cytochrome
P450
XIX
(
aromatase),
which
converts
testosterone
to
estradiol.
Peroxisome
proliferators
are
known
to
induce
 ­
oxidation
and
cytochrome
P­
450
monooxygenases
by
binding
to
the
peroxisome
proliferation
activation
receptor
 
(
PPAR
 ;
a
subfamily
of
steroid
hormone
receptors).
It
is
believed
that
APFO
induces
cytochrome
P450
XIX
(
aromatase)
by
binding
to
and
activating
the
PPAR .

Although
significant
increases
in
LH
were
not
seen
in
Sprague­
Dawley
rats
after
treatment
of
APFO
in
the
14
day­
studies,
it
appears
that
increase
in
LH
levels
cannot
be
ruled
out
to
be
involved
(
in
addition
to
increased
estradiol
level)
in
the
induction
of
LCT
by
APFO.
In
these
studies,
significant
increase
in
hepatic
aromatase
(
which
converts
testosterone
to
estradiol)
activities
associated
with
decreased
serum
testosterone
levels
and
increased
estradiol
levels
were
observed
in
the
treated
rats.
Testosterone
is
synthesized
and
secreted
by
the
Leydig
cells,
and
is
regulated
by
LH;
testosterone
and
LH
form
a
closed­
loop
feedback
system
in
the
HPT
axis.
In
order
to
maintain
adequate
testosterone
plasma
levels,
reduced
testosterone
levels
(
caused
by
increased
aromatase
activity)
are
expected
to
lead
to
increased
LH
levels
through
the
negative
feedback
mechanism.
It
has
been
pointed
out
that
increases
in
LH
may
not
always
be
seen
in
all
studies
of
chemicals
for
which
the
proposed
mode
of
action
calls
for
elevated
LH,
and
that
compensation
may
have
occurred
to
restore
homeostasis
and
inappropriate
timing
of
sampling
are
some
of
the
explanations
for
failing
to
detect
changes
in
LH
levels
(
Clegg
et
al.,
1997).

3.8.2.3
Mammary
Gland
Tumors
Estradiol
has
also
been
shown
to
stimulate
the
secretion
of
TGF
 
by
mammary
epithelial
cells
and
the
overexpression
of
TGF
 
has
been
suggested
as
one
possible
factor
in
producing
sustained
cell
proliferation
of
mammary
tumor
cells
and
the
subsequent
development
of
neoplasia
(
Liu
et
al.,
1987).
Hence,
it
is
possible
that
the
APFO­
induced
elevation
of
estradiol
levels
may
also
be
responsible
for
the
development
of
mammary
fibroadenomas
in
Sprague
Dawley
rats
in
addition
to
LCT
(
discussed
above).
In
fact,
this
is
consistent
with
the
mechanism
by
which
spontaneous
mammary
neoplasms
were
developed
in
aging
female
Sprague
Dawley
rats.
It
has
been
demonstrated
that
the
early
appearance
and
high
spontaneous
incidence
of
mammary
gland
tumors
in
untreated,
aging
female
Sprague­
Dawley
rats
is
due
to
increased
exposure
to
endogenous
estrogen
and
prolactin
as
a
result
of
an
accelerating
effect
on
normal,
age­
related
perturbations
of
the
estrous
cycle
in
this
strain
of
rat
(
Cutts
and
Noble,
1964;
Chapin
et
al.,
1996).

3.8.2.4
Pancreatic
Tumors
The
mechanism
by
which
APFO
induced
pancreatic
acinar
cell
tumors
is
unknown.
A
number
of
other
peroxisome
proliferators
also
produce
pancreatic
acinar
cell
tumors
in
rats.
Available
data
suggest
that
the
pancreatic
acinar
cell
tumors
are
related
to
an
increase
in
serum
cholecystokinin
(
CCK)
level
secondary
to
hepatic
cholestasis
(
Cook
et
al.,
1994;
Obourn
et
al.,
1997).
CCK
is
a
78
growth
factor
that
has
been
shown
to
stimulate
normal,
adaptive,
and
neoplastic
growth
of
pancreatic
acinar
cells
in
rats
(
Longnecker,
1987).
However,
data
on
the
role
of
CCK
in
pancreatic
tumor
formation
are
conflicting.

3.9
Immunotoxicology
Studies
in
Animals
Four
immunotoxicity
studies
of
PFOA
have
been
conducted
in
mice.
The
first
of
these
studies
was
a
feeding
study
in
mice
(
Yang
et
al.
2000).
For
investigation
of
the
effects
of
perfluorooctanoic
acid
(
PFOA),
and
other
peroxisome
proliferators,
on
lymphoid
organs,
0.02
%
PFOA
was
administered
to
male
C57Bl/
6
mice
in
the
diet
for
2,
5,
7,
or
10
days.
At
the
end
of
the
feeding
period,
mice
were
sacrificed
and
the
liver,
spleen,
and
thymus
were
dissected
out
and
weighed.
The
effect
of
PFOA
administration
on
the
cellularity,
cell
surface
phenotype,
and
cell
cycle
of
thymocytes
and
splenocytes
was
determined.
In
addition,
effects
of
exposure
of
thymocytes
and
splenocytes
to
PFOA
in
vitro
were
examined.

The
results
showed
that
administration
of
0.02%
PFOA
for
2,
5,
7,
or
10
days
resulted
in
a
significant
increase,
relative
to
controls,
in
liver
weight,
even
at
the
earliest
time
point.
Also,
a
decrease
in
body
weight
caused
by
PFOA
administration
was
observed.
Subsequently,
by
the
day
5
administration
period,
significant
decreases
in
thymus
and
spleen
weight
were
detected.
After
administration
of
0.02%
PFOA
for
7
days,
significant
decreases
(
85%
and
80%,
respectively)
in
the
total
number
of
thymocytes
and
splenocytes
were
observed.
The
results
also
showed
that
the
number
of
thymocytes
expressing
both
CD4
and
CD8
decreased
by
95%;
the
number
expressing
both
CD4
and
CD8
decreased
by
57%;
and
the
number
expressing
either
CD4
or
CD8
decreased
by
64%
and
72%,
respectively.
For
the
splenocytes,
both
T
cells
(
CD3)
and
B
cells
(
CD19)
decreased
by
75%
and
86%,
respectively.
Also,
significant
decreases
in
both
CD4
helper
and
CD8
cytotoxic
splenic
T
cells
were
observed.
Upon
administration
of
0.02%
PFOA
to
mice
for
7
days,
thymocyte
proliferation
was
also
inhibited,
as
detected
by
cell
cycle
flow
cytometry
analyses.
In
vitro
studies
showed
that
there
was
spontaneous
apoptosis
occurring
in
splenocytes
and
thymocytes
after
8
or
24
hours
of
culturing
in
the
presence
of
varying
concentrations
(
50,
100,
or
200
M)
of
PFOA.
However,
PFOA
did
not
significantly
alter
the
cell
cycle
under
these
conditions.

In
order
to
study
mechanism
(
Yang
et
al.
2001)
,
another
mouse
feeding
study
was
performed.
In
order
to
examine
the
dose
dependency
of
the
effects,
C57Bl/
6
mice
received
diets
consisting
of
0.001%­
0.05%
PFOA
(
w/
w)
for
10
days.
For
examining
the
time­
course,
a
diet
containing
0.02%
PFOA
was
given
for
2,
5,
7
or
10
days.
Effects
of
withdrawal
of
PFOA
were
also
studied.

The
results
showed
that,
at
higher
doses,
a
significant
decrease,
relative
to
controls,
in
body
weight
was
observed,
although
no
other
apparent
signs
of
toxicity
such
as
sores,
lethargy,
and
poor
grooming
were
noticed.
However,
a
significant
decrease
in
total
water
intake
was
observed.
Mice
receiving
dietary
PFOA
for
10
days
experienced
significant
increases
in
liver
weight
and
peroxisome
proliferation,
as
measured
by
induction
of
acyl­
CoA
oxidase
with
lauroyl­
CoA
or
palmitoyl­
CoA
as
substrate.
These
increases
started
at
the
lowest
dose
and
79
reached
their
maximal
values
at
a
dose
of
0.003­
0.01%.
In
contrast,
the
weight
decreases
of
the
spleen
and
thymus
began
at
a
higher
dose
(
0.01%)
with
no
maximum
reached
with
the
doses
given.
The
time
course
studies
showed
that
increased
liver
weights
and
peroxisome
proliferation
were
evident
at
the
earliest
time
point
examined.
In
contrast,
significant
thymus
and
spleen
weight
decreases
required
PFOA
administration
for
a
period
of
at
least
5
days,
following
which
the
spleen
weight
remained
constant
while
the
thymus
weight
continued
to
decrease.
However,
upon
prolonged
treatment
for
one
month,
no
further
decreases
in
thymus
and
spleen
weights
were
observed.
In
another
set
of
experiments,
animals
received
0.02%
PFOA
for
7
days,
and
then
they
received
normal
chow
for
a
period
of
10
days.
These
recovery
experiments
showed
that
the
animals
rapidly
recovered
the
body
weight
the
second
day
after
withdrawal
of
PFOA.
However,
the
liver
weight
did
not
return
to
normal
even
after
10
days
of
recovery.
Thymus
recovery
started
on
day
2
and
was
completed
by
day
10.
The
spleen
weights
returned
to
normal
by
day
2
post­
withdrawal.
In
addition,
the
changes
in
thymus
and
spleen
weight
upon
PFOA
treatment
and
withdrawal
paralleled
the
changes
in
total
thymocyte
and
splenocyte
counts.
Furthermore,
flow
cytometry
cell
cycle
experiments
showed
that
the
decrease
in
thymocyte
number
caused
by
PFOA
treatment
is
due
mainly
to
inhibition
of
thymocyte
proliferation.
In
contrast,
PFOA
treatment
caused
no
changes
in
the
cell
cycle
of
splenocytes.

A
third
feeding
study
(
Yang
et
al.
2002a)
was
designed
to
examine
the
possible
involvement
of
the
peroxisome
proliferator­
activated
receptor
alpha
(
PPAR 
in
the
immunomodulation
exerted
by
PFOA.
This
study
made
use
of
transgenic
PPAR 
null
mice,
which
are
homozygous
with
regards
to
a
functional
mutation
in
the
PPAR 
gene.
These
mice
do
not
exhibit
peroxisome
proliferation
or
hepatomegaly
and
hepatocarcinogenesis
even
after
exposure
to
peroxisome
proliferators.
These
mice
were
fed
a
diet
consisting
of
0.02%
PFOA
(
w/
w)
for
7
days.
At
the
end
of
the
feeding
period,
mice
were
sacrificed
and
the
liver,
spleen,
and
thymus
were
removed
and
weighed.
The
effect
of
PFOA
on
peroxisome
proliferation,
cell
cycle,
and
lymphoproliferation
was
ascertained.

The
results
showed
that,
in
contrast
to
wild­
type
mice,
feeding
PPAR 
null
mice
PFOA
resulted
in
no
significant
decrease
in
body
weight.
However,
increases
in
liver
weight
were
still
seen
in
PPAR 
null
mice,
suggesting
that
this
is
not
a
PPAR 
dependent
process.
As
expected,
peroxisome
proliferation,
as
measured
by
fatty
acid
­
oxidation,
was
totally
lacking
in
PPAR 
null
mice.
Also
in
contrast
to
wild
type
mice,
feeding
PPAR(
null
mice
PFOA
resulted
in
no
significant
decrease
in
the
weight
of
the
spleen
or
the
number
of
splenocytes.
At
the
same
time,
the
decrease
in
weight
and
cellularity
of
the
thymus
was
attenuated,
but
not
totally
eliminated
in
the
PPAR(
null
mice.
In
addition,
the
decreases
in
the
size
of
the
CD4+
CD8+
population
of
thymus
cells
and
the
number
of
thymus
cells
in
the
S
and
G2/
M
phases
of
the
cell
cycle,
which
reflects
inhibition
of
proliferation,
observed
in
wild
type
mice
administered
PFOA
were
much
less
extensive
in
PPAR(
null
mice.
Finally,
in
contrast
to
wild
type
mice,
PFOA
treatment
caused
no
significant
change
in
splenocyte
proliferation
in
response
to
mitogens
in
PPAR(
null
mice.

A
fourth
feeding
study
(
Yang
et
al.
2002b)
was
designed
to
examine
the
effects
of
PFOA
on
specific
humoral
immune
responses
in
mice.
For
this
study,
0.02
%
PFOA
was
administered
to
80
male
C57Bl/
6
mice
for
10
days.
Then
the
animals
were
examined,
via
plaque
forming
cell
(
PFC)
and
serum
antibody
assays,
for
their
ability
to
generate
an
immune
response
to
horse
red
blood
cells
(
HRBCs).
Ex
vivo
and
in
vitro
splenic
lymphocyte
proliferation
assays
were
also
performed.

The
results
showed
that
mice
fed
normal
chow
responded
to
challenge
with
HRBCs
with
a
strong
humoral
response,
as
measured
by
the
PFC
assay.
In
contrast,
mice
fed
with
PFOA
responded
to
HRBC
immunization
with
no
increase
in
HRBC­
specific
PFCs,
relative
to
unimmunized
controls.
However,
in
experiments
where
PFOA­
treated
mice
received
normal
chow
following
HRBC
immunization,
there
was
a
significant
recovery
of
the
numbers
of
specific
PFCs
stimulated.
The
suppression
of
the
humoral
immune
response
by
PFOA
was
confirmed
by
analysis
of
the
serum
anti­
HRBC
response.
In
ex
vivo
experiments,
splenocytes
isolated
from
control
mice
responded
to
both
ConA
and
LPS
with
lymphocyte
proliferation,
as
measured
by
thymidine
incorporation.
However,
treating
mice
with
PFOA
(
0.02%
for
7
days)
attenuated
the
proliferation.
In
a
set
of
in
vitro
experiments,
PFOA
(
1­
200
M)
added
to
the
culture
medium
of
splenocytes
cultured
from
untreated
mice
did
not
cause
an
alteration
of
lymphocyte
proliferation
in
response
to
LPS
or
ConA.

4.0
Hazards
to
the
Environment
4.1
Introduction
The
aquatic
toxicity
and
hazard
of
APFO
to
aquatic
organisms
was
assessed.
This
task
was
made
more
difficult
by
several
problems
discussed
below.
These
problems
complicated
the
task
of
determining
if
the
ecotoxicity
tests
were
valid
and
could
be
used
in
the
assessment.
Furthermore,
these
problems
limited
the
confidence
that
could
be
placed
on
the
toxicity
test
values,
and
thus
in
turn
lowered
the
confidence
of
conclusions
that
could
be
drawn
in
assessing
the
inherent
toxicity
and
hazard
of
APFO
to
aquatic
organisms.

1)
A
variety
of
different
APFOs
with
varying
designations
and
lot
numbers
were
tested.
Generally,
the
ammonium
salt
or
the
tetrabutylammonium
salt
was
tested.
The
exact
composition
and
identification
of
impurities,
which
may
affect
toxicity,
in
each
lot
number
used
is
not
known.

2)
A
variety
of
testing
laboratories
conducted
the
APFO
toxicity
studies
over
a
period
of
time
from
approximately
1974­
1996.
This
situation
served
to
increase
overall
test
variability
and
thus
made
inter­
laboratory
comparisons
more
difficult.

3)
Purity
of
the
tested
material,
or
percent
test
material
and
percent
other
material(
s),
was
a
major
concern.
Purity
was
not
sufficiently
characterized
in
these
tests.
In
some
tests
it
appeared
that
100%
test
chemical
was
used;
in
others
a
chemical
of
lesser
purity
(
approximately
85%)
was
used.
Purity
of
test
material
does
affect
toxicity
and
should
be
taken
into
account
when
possible,
by
expressing
toxicity
on
the
same
purity
basis.

4)
Water,
an
isopropanol
solvent,
or
a
combination
of
both
were
used
with
the
test
material
in
81
many
of
the
toxicity
tests,
for
no
obvious
indicated
reason.
Solvents
are
mixed
with
the
test
material
to
make
it
miscible
with
the
test
dilution
water
before
the
test
is
begun.
Solvents
are
used
in
tests
where
the
concentrations
of
the
test
material
are
extremely
low
and
a
very
small
amount
of
test
material
must
be
added
to
the
test
chambers.
It
was
not
clear
from
the
summaries
of
these
studies
why
a
solvent
was
used
or
was
even
found
to
be
necessary.
In
fact,
3M
summarized
each
test
and
stated
"
Data
may
not
accurately
relate
toxicity
of
the
test
sample
with
that
of
the
test
substance."
Thus,
in
those
tests
where
100%
test
material
was
not
used,
the
toxicity
values
had
to
be
adjusted
to
take
into
account
the
percent
solvent(
s),
and
to
express
the
values
on
a
100%
test
chemical
basis,
so
that
the
tests
could
be
compared.

5)
In
all
these
toxicity
tests
only
nominal
test
chemical
concentrations
were
used.
Measured
test
chemical
concentrations
are
instead
always
recommended
so
that
one
can
accurately
determine
the
actual
test
chemical
concentration
to
which
the
test
organisms
are
exposed.
If
it
is
determined
that
the
nominal
concentrations
are
only,
for
example
50%
of
the
measured
concentrations,
the
toxicity
values
will
have
to
accordingly
be
adjusted
by
50%.
Analytical
measurements
of
chemical
concentration
should
have
been
taken
or
made
available.
Then,
recovery
rates
could
have
been
determined,
and
physicochemical
processes
(
e.
g.,
hydrolysis,
volatility)
that
might
lower
the
actual
concentrations
to
which
the
test
organisms
were
exposed
could
have
been
taken
into
account.
Nominals
may
be
used
when
measured
concentrations
are
taken
and
the
relationship
of
both
is
known.

In
order
to
proceed
with
any
sort
of
environmental
hazard
review
it
was
necessary
to
ignore
these
test
limitations
and
to
assume
that
the
nominal
concentrations
were
an
"
adequate"
expression
of
the
measured
test
chemical
concentrations.
Criteria
for
assessing
degree
of
acute
toxicity
are
based
on
well­
established
values
(
low
is
>
100
mg/
L;
medium
or
moderate
is
>
1<
100
mg/
L;
high
is
<
1
mg/
L).

4.2
Acute
Toxicity
to
Freshwater
Species
Several
species
were
tested
to
assess
the
acute
toxicity
of
APFO;
these
included
the
fathead
minnow
(
Pimephales
promelas),
bluegill
sunfish
(
Lepomis
machrochirus),
water
flea
(
Daphnia
magna),
and
a
green
alga
(
Selenastrum
capricornutum).
The
toxicity
test
endpoints
have
been
adjusted
to
100%
test
chemical
and
test
results
are
presented
in
Tables
9
(
organized
by
test
substance)
and
10
(
organized
by
test
species).
Each
value
is
related
to
a
testing
facility
and
reference.

Twelve
tests
were
conducted
with
fathead
minnows;
96­
h
LC50
values
(
based
on
mortality)
ranged
from
70
to
843
mg/
L.
It
is
unclear
why
this
range
is
so
wide.
Assuming
these
studies
are
valid,
and
due
to
the
limitations
discussed
above,
these
toxicity
values
indicate
low
toxicity.
The
two
acute
values
for
bluegill
sunfish
also
indicate
low
toxicity
(
96­
h
LC50s
of
>
420,
and
569
mg/
L).

Nine
acute
tests
were
conducted
with
daphnids
and
48­
h
EC50
values
(
based
on
immobilization)
ranged
from
39
to
>
1000
mg/
L.
The
lower
values
are
indicative
of
moderate
toxicity,
but
the
82
wide
range
makes
interpretation
difficult.

Seven
tests
were
conducted
with
green
algae;
96­
h
EC50
values
(
based
on
growth
rate,
cell
density,
cell
counts,
and
dry
weights)
ranged
from
1.2
to
>
666
mg/
L
(
the
Er50
cell
density
value
of
1,000
mg/
L
is
excluded
from
this
discussion).
The
lower
value
indicates
high
to
moderate
toxicity,
based
on
the
acute
criteria.
The
lower
value
would
also
be
indicative
of
moderate
toxicity,
based
on
the
chronic
moderate
criterion
(.
0.1<
10
mg/
L).
A
14­
d
EC50
value
of
43
mg/
L,
based
on
cell
counts,
for
green
algae
was
also
calculated
in
one
study.
This
is
indicative
of
low
chronic
toxicity,
based
on
the
chronic
criterion
(
10
mg/
L).
Green
algae
appeared
to
be
the
most
sensitive
test
species
in
the
44%
APFO
test
sample,
daphnids
were
the
next
most
sensitive,
and
fathead
minnows
were
the
least
sensitive.
83
Table
9.
Summary
of
Acute
Ecological
Toxicity
Data
for
APFO
(
grouped
by
test
substance)
Test
Organism
Duration
Value
(
mg/
L)*
Reference
Test
Sample:
APFO
ammonium
salt
96­
h
LC50
70
3M
Company,
1974a
96­
h
LC50
766
3M
Company,
1980a
96­
h
LC50
301
3M
Company,
1987c
Fathead
minnow
(
Pimephales
promelas)

96­
h
LC50
740
Ward
et
al.,,
1995
96­
h
LC50
>
420
3M
Company,
1978
Bluegill
sunfish
(
Lepomis
machrochirus)
96­
h
LC50
569
3M
Company,
1978
48­
h
EC50
126
3M
Environmental
Laboratory,
1982
48­
h
EC50
>
1000
3M
Environmental
Laboratory,
1982
48­
h
EC50
221
3M
Company,
1987b
Water
flea
(
Daphnia
magna)

48­
h
EC50
720
Ward
et
al.,
1995
96­
h
EC50
310
Ward
et
al.,
1995
Green
algae
(
Selenastrum
capricornutum)
96­
h
EC50
1000
Ward
et
al.,
1995
30­
min
EC50
870
3M
Company,
1987a
Bacteria
(
Photobacterium
phosphoreum)
30­
min
EC50
730
3M
Environmental
Laboratory,
1996a
7­
min
NOEC
1000
3M
Company,
1980b
Activated
sludge
30­
min
EC50
>
1000
3M
Company,
1987d
Test
Sample:
APFO
96­
h
LC50
440
3M
Company,
1974b
Fathead
minnow
(
Pimephales
promelas)
96­
h
LC50
843
3M
Company,
1985
Test
Sample:
APFO
ammonium
salt
in
50%
water
96­
h
LC50
>
500
EnviroSystems,
Inc.,
1990a
Fathead
minnow
(
Pimephales
promelas)

96­
h
NOEC
500
EnviroSystems,
Inc.,
1990a
Water
flea
(
Daphnia
magna)
48­
h
EC50
292
EnviroSystems,
Inc.,
1990b
Bacteria
(
Photobacterium
phosphoreum)
30­
min
EC50
>
500
3M
Environmental
Laboratory,
1990a
Test
Sample:
APFO
ammonium
salt
in
50%
water,
continued
Activated
sludge
30­
min
EC50
>
500
3M
Environmental
Laboratory,
1990b
84
Test
Sample:
APFO
ammonium
salt
in
80%
water
Fathead
minnow
(
Pimephales
promelas)
96­
h
LC50
494
Ward
et
al.,
1996a
Water
flea
(
Daphnia
magna)
48­
h
EC50
240
Ward
et
al.,
1996c
96­
h
EC50
396
Ward
et
al.,
1996b
Green
algae
(
Selenastrum
capricornutum)
96­
h
EC50
>
666
Ward
et
al.,
1996b
30
min
EC50
630
3M
Environmental
Laboratory,
1996b
Bacteria
(
Photobacterium
phosphoreum)
30
min
EC50
390
3M
Environmental
Laboratory,
1996c
Activated
sludge
30­
min
EC50
>
664
3M
Environmental
Laboratory,
1996d
Test
Sample:
APFO
in
50%
isopropanol
Fathead
minnow
(
Pimephales
promelas)
96h
LC50
140
T.
R.
Wilbury
Laboratories,
Inc.,
1996a
Water
flea
(
Daphnia
magna)
48­
h
EC50
360
T.
R.
Wilbury
Laboratories,
Inc.,
1996b
Green
algae
(
Selenastrum
capricornutum)
96­
h
EC50
90
T.
R.
Wilbury
Laboratories,
Inc.,
1995
Test
Sample:
APFO
(
44%)
in
27.9%
water
and
27.2%
isopropanol
Fathead
minnow
(
Pimephales
promelas)
96­
h
EC50
391
T.
R.
Wilbury
Laboratories,
Inc.,
1995
Fathead
minnow
(
Pimephales
promelas)
96­
h
EC50
422
T.
R.
Wilbury
Laboratories,
Inc.,
1995
Test
Sample:
APFO
(
44%)
in
27.9%
water
and
27.2%
isopropanol
Water
flea
(
Daphnia
magna)
48­
h
EC50
41
Ward
et
al.,
1995
Water
flea
(
Daphnia
magna)
48­
h
EC50
39
Ward
et
al.,
1995
Green
algae
(
Selenastrum
capricornutum)
96­
h
EC50
2.1
Ward
et
al.,
1995
Green
algae
(
Selenastrum
capricornutum)
96­
h
EC50
3.6
Ward
et
al.,
1995
Green
algae
(
Selenastrum
capricornutum)
96­
h
EC50
1.2
Ward
et
al.,
1995
*
Values
were
adjusted
to
represent
100%
active
ingredient.
AThese
values
may
be
inconsistent
due
to
different
diets
tested.
85
Table
10.
Summary
of
Ecological
Toxicity
Data
for
APFO
(
grouped
by
species)
Test
Organism
Duration
Value
(
mg/
L)
Reference
96­
h
LC50
70B
3M
Company,
1974a
96­
h
LC50
766B
3M
Company,
1980a
96­
h
LC50
301B
3M
Company,
1987c
96­
h
LC50
440C
3M
Company,
1974b
96­
h
LC50
843C
3M
Company,
1985
96­
h
LC50
>
500D
EnviroSystems,
Inc.,
1990a
96­
h
NOEC
500D
EnviroSystems,
Inc.,
1990a
96­
h
LC50
494
Ward
et
al.,
1996a
96h
LC50
140F
T.
R.
Wilbury
Laboratories,
Inc.,
1996a
30­
day
NOAEL
>
100B
EG&
G
Bionomics
Aquatic
Toxicology
Laboratory,
1978
96­
h
EC50
391G
T.
R.
Wilbury
Laboratories,
Inc.,
1995
Fathead
minnow
(
Pimephales
promelas)

96­
h
EC50
422G
T.
R.
Wilbury
Laboratories,
Inc.,
1995
96­
h
LC50
>
420B
3M
Company,
1978
Bluegill
sunfish
(
Lepomis
machrochirus)
96­
h
LC50
569B
3M
Company,
1978
48­
h
EC50
126AB
3M
Environmental
Laboratory,
1982
48­
h
EC50
>
1000AB
3M
Environmental
Laboratory,
1982
48­
h
EC50
221B
3M
Company,
1987b
48­
h
EC50
292D
EnviroSystems,
Inc.,
1990b
48­
h
EC50
240
Ward
et
al.,
1996c
48­
h
EC50
360F
T.
R.
Wilbury
Laboratories,
Inc.,
1996b
21­
day
IC50
43B
3M
Company,
1984
21­
day
NOEC
22B
3M
Company,
1984
21­
day
NOEC
22B
3M
Company,
1984
48­
h
EC50
41G
Ward
et
al.,
1995
Water
flea
(
Daphnia
magna)

48­
h
EC50
39G
Ward
et
al.,
1995
86
Table
10.
Summary
of
Ecological
Toxicity
Data
for
APFO
(
grouped
by
species)
Test
Organism
Duration
Value
(
mg/
L)
Reference
96­
h
EC50
396
Ward
et
al.,
1996b
96­
h
EC50
>
666E
Ward
et
al.,
1996b
96­
h
EC50
90F
T.
R.
Wilbury
Laboratories,
Inc.,
1995
14­
day
EC50
43B
Elnabarawy,
1981
96­
h
EC50
2.1G
Ward
et
al.,
1995
96­
h
EC50
3.6G
Ward
et
al.,
1995
Green
algae
(
Selenastrum
capricornutum)

96­
h
EC50
1.2G
Ward
et
al.,
1995
30­
min
EC50
870B
3M
Company,
1987a
30­
min
EC50
730B
3M
Environmental
Laboratory,
1996a
30­
min
EC50
>
500D
3M
Environmental
Laboratory,
1990a
30
min
EC50
630
3M
Environmental
Laboratory,
1996b
Bacteria
(
Photobacterium
phosphoreum)

30
min
EC50
390
3M
Environmental
Laboratory,
1996c
7­
min
NOEC
1000B
3M
Company,
1980b
30­
min
EC50
>
1000B
3M
Company,
1987d
30­
min
EC50
>
500D
3M
Environmental
Laboratory,
1990b
Activated
sludge
30­
min
EC50
>
664E
3M
Environmental
Laboratory,
1996d
*
Values
were
adjusted
to
represent
100%
active
ingredient.
AThese
values
may
be
inconsistent
due
to
different
diets
tested.
BTested
substance
was
APFO
ammonium
salt.
CTested
substance
was
APFO
DTested
substance
was
APFO
ammonium
salt
in
50%
water.
ETested
substance
was
APFO
ammonium
salt
in
80%
water.
FTested
substance
was
APFO
in
50%
isopropanol.
GTest
Sample:
APFO
(
44%)
in
27.9%
water
and
27.2%
isopropanol
87
5.0
References
3M
Company.
1976a.
Primary
Eye
Irritation
Study­
Rabbits.

3M
Company.
1976b.
Acute
Oral
Toxicity
in
Rats­
T­
1585.

3M
Company.
1977.
Ready
Biodegradation
of
FC­
143
(
BOD/
COD/
TOC).
Environmental
Laboratory.
St.
Paul,
MN.

3M
Company
International
Research
and
Development
Corporation.
1978b.
Fluorad
Fluorochemical
FC­
143
Acute
Oral
Toxicity
(
LD50)
Study
in
Rats.
Study
No.
137­
091,

3M
Company.
1979.
Technical
Report
Summary
­
Final
Comprehensive
Report:
FC­
143.
(
USEPA
AR­
226
528)

3M
Company.
1980c.
Ready
Biodegradation
of
FC­
143(
BOD/
COD).
Lab
Request
No.
5625S.
Environmental
Laboratory.
St.
Paul,
MN.

3M
Company.
1981.
Water,
Acetone
and
Toluene
Solubility
Estimates.
Environmental
Laboratory.
St.
Paul,
MN.

3M
Company.
1984.
Chronic
toxicity
to
freshwater
invertebrates.

3M
Company.
1985a.
96­
hour
acute
static
toxicity
to
fathead
minnow
 
FX­
1001.
Environmental
Laboratory,
St.
Paul,
MN.
Lab
Request
Number
C1006.
February
2.

3M
Company.
1985b.
Ready
Biodegradation
of
FX­
1001
(
BOD/
COD).
Lab
Request
No.
C1006.
Environmental
Laboratory.
St.
Paul,
MN.
February
14.

3M
Company.
1987.
Activated
Sludge
Respiration
Inhibition
Test.
Environmental
Laboratory;
Lab
Request
Number
E1282.
St.
Paul,
MN.

3M
Company.
1999a.
"
The
Science
of
Organic
Fluorochemistry"
and
"
Perfluorooctane
sulfonate:
Current
summary
of
human
sera,
health,
and
toxicology
data".
February
5,
1999.
(
8EHQ­
0299­
373).

3M
Company.
1999b.
8EHQ­
0699­
373.
Supplement.
May
26,
1999.

3M
Company.
2000a.
Voluntary
Use
and
Exposure
Information
Profile
for
Perfluorooctanesulfonic
Acid
and
Various
Salt
Forms.
3M
Company
submission
to
USEPA,
dated
April
27,
2000.

3M
Company.
2000b.
Voluntary
Use
and
Exposure
Information
Profile
for
Perfluorooctanoic
Acid
and
Salts.
3M
Company
submission
to
USEPA,
dated
June
8,
2000.
88
3M
Company,
2000c.
About
3M
Worldwide:
3M
Phasing
Out
Some
of
Its
Specialty
Materials.
http://
www.
3M.
com/
about3m/
worldwide/
release.
html.

3M
Company.
2001a.
Environmental
Monitoring
­
Multi­
City
Study,
3M
Environmental
Laboratory,
June
25.
In
U.
S.
EPA
Administrative
Record
AR226­
1030A.

3M
Company,
2002.
Submission
dated
September
17,
2002
to
USEPA,
AR
226.

3M
Environmental
Laboratory.
1990a.
Microbics
Microtox
®
Toxicity
Test.
St.
Paul,
Minnesota.
Lab
request
number
G2882.

3M
Environmental
Laboratory.
1990b.
Activated
Sludge
Respiration
Inhibition.
St.
Paul,
Minnesota.
Lab
Request
number
G2882.

3M
Environmental
Laboratory.
1993.
Impinger
Studies
of
Volatility
of
FC­
95
and
FC­
143.
St.
Paul,
MN.
3M
Laboratories.
3M
Lab
Request
Number
L3306.

3M
Environmental
Laboratory.
1996a.
Microbics
Microtox
®
Toxicity
Test
of
FC­
143.
Lab
Request
number
P1626.
St.
Paul,
Minnesota.

3M
Environmental
Laboratory.
1996b.
Microbics
Microtox
®
Toxicity
Test
of
FC­
118.
Lab
Request
number
P1626.
St.
Paul,
Minnesota.

3M
Environmental
Laboratory.
1996c.
Microbics
Microtox
®
Toxicity
Test
of
FC­
1015­
X.
Lab
Request
number
P1626.
St.
Paul,
Minnesota.

3M
Environmental
Laboratory.
2001a.
Hydrolysis
Reactions
of
Perfluorooctanoic
Acid
(
PFOA).
Lab
Request
Number
E00­
1851.
March
30.

3M
Environmental
Laboratory.
2001b.
Characterization
Study
of
PFOA
(
lot
#
332),
Primary
Standard
 
Test
Control
Reference
#
TCR­
99030­
030.
Phase:
Solubility
Determination.
3M
Laboratories,
St.
Paul,
MN.

3M
Environmental
Laboratory
2001c.
26­
Week
Capsule
Toxicity
Study
with
Ammonium
Perfluorooctanoate
(
APFO)
in
Cynomolgus
Monkeys
Amended
Analytical
Laboratory
Report
Title
Determination
of
the
Presence
and
Concentration
of
Perfluorooctanoate
Fluorochemical
in
Liver,
Serum,
Urine
and
Feces
Samples..
Performing
Laboratories,
Liver,
Serum
and
Urine
Analyses,
3M
Environmental
Laboratories,
St.
Paul,
MN,
Feces
Analyses
Centre
Analytical
Laboratories,
Inc.
State
College,
PA.
Analytical
Phase
Completion
Date,
June
11,
2001,
408
pp.

Abdellatif,
A.
G.
Preat,
V.,
Taper,
H.
S.
and
Roberfroid,
M.
1991.
The
modulation
of
rat
liver
carcinogenesis
by
perfluorooctanoic
acid,
a
peroxisome
proliferator.
Toxicol.
Appl.
Pharmacol.
111:
530­
537.
89
Alexander,
B.
H.
2001a.
Mortality
study
of
workers
employed
at
the
3M
Cottage
Grove
facility.
Final
Report.
Division
of
Environmental
and
Occupational
Health,
School
of
Public
Health,
University
of
Minnesota,
April
26,
2001.

Alexander,
B.
H.
2001b.
Mortality
study
of
workers
employed
at
the
3M
Decatur
facility.
Final
Report.
Division
of
Environmental
and
Occupational
Health,
School
of
Public
Health,
University
of
Minnesota,
April
26,
2001.

Behar,
B.;
Stein,
G.
Science
1966,
Vol.
154,
p.
1012.

Beach,
S.
1995a.
Inhibitory
Effect
of
L­
13492
to
Microbics'
MicrotoxTM
Toxicity
Analyzer
System.
3M
Company,
Environmental
Laboratory,
St
Paul,
MN,
Lab
Request
number
N2169.
July
26.

Beach,
S.
1995b.
Inhibitory
Effect
of
L­
13492
on
Activated
Sludge.
3M
Company
Environmental
Laboratory,
St.
Paul,
MN,
Lab
Request
N2169,
July
26.

Biegel,
L.
B.,
Liu,
R..
C.
M.,
Hurtt,
M.
E.
and
Cook,
J.
C.
1995.
Effects
of
ammonium
perfluorooctanate
on
Leydig
cell
function:,
,
and
ex
vivo
studies.
Toxicol.
Appl.
Pharmacol.
134:
18­
25.

Biegel,
L.
B.,
Hurtt,
M.
E.,
Frame,
S.
R.,
O'Connor,
J.
C.,
and
J.
C.
Cook.
2001.
Mechanisms
of
extrahepatic
tumor
induction
by
peroxisome
proliferators
in
male
CD
rats.
Toxicological
Sciences.
60:
44­
55.

Bio/
dynamics
Inc.
1979.
An
Acute
Inhalation
Study
of
T­
2305
CoC
in
the
Rat.
3M
Company,
St.
Paul,
MN.
Project
No.
78­
7184.

Biosearch,
Inc.
1976.
Primary
Eye
Irritation
Study
 
Rabbits.
Philadelphia,
PA.
3M
Company.
St.
Paul,
MN.

Boeri,
R.,
Magazu,
J.,
Ward,
T.
1995a.
Acute
toxicity
of
L­
13492
to
the
Daphnid,
Daphnia
Magna.
T.
R.
Wilbury
Laboratories,
Inc.,
3M
Company
Lab
Request
number
N2332,
July
13.

Boeri,
R.,
Magazu,
J.,
Ward,
T.
1995b.
Acute
Toxicity
of
L­
13492
to
the
Fathead
Minnow,
Pimephales
promelas.
T.
R.
Wilbury
Laboratories,
Inc.,
3M
Company
Lab
Request
number
N2332,
July
13.

Boeri,
R.,
Magazu,
J.,
Ward,
T.
1995c.
Growth
and
reproduction
Toxicity
Test
with
L­
13492
and
the
Freshwater
Alga,
Selenastrum
capricornutum.
T.
R.
Wilbury
Laboratories,
Inc.,
3M
Company
Lab
Request
number
N2332.
August
3.

Boeri,
R.,
Kowalski,
P.,
Ward,
T.
1995d.
Acute
Toxicity
of
N2803­
2
to
the
Fathead
Minnow,
90
Pimephales
promelas.
T.
R.
Wilbury
Laboratories,
Inc.,
3M
Company,
Lab
Request
number
N2803­
2.
November
16.

Boeri,
R.,
Kowalski,
P.,
Ward,
T.
1995e.
Acute
Toxicity
of
N2803­
4
to
the
Fathead
Minnow,
Pimephales
promelas.
T.
R.
Wilbury
Laboratories,
Inc.,
3M
Company
Lab
Request
number
N2803­
4.
November
21.

Boeri,
R.,
Kowalski,
P.,
Ward,
T.
1995f.
Growth
and
Reproduction
Toxicity
Test
with
N2803­
2
and
the
Freshwater
Algae,
Selenastrum
capicornutum.
T.
R.
Wilbury
Laboratories,
Inc.,
3M
Company
Lab
Request
number
N2803­
2.
November
28.

Boeri,
R.,
Kowalski,
P.,
Ward,
T.
1996a.
Growth
and
Reproduction
Toxicity
Test
with
N2803­
4
and
the
Freshwater
Alga,
Selenastrum
capicornutum.
T.
R.
Wilbury
Laboratories,
Inc.,
3M
Company
Lab
Request
number
N2803­
4.
March
7.

Boeri,
R.,
Kowalski,
P.,
Ward,
T.
1996b.
Acute
Toxicity
of
N2803­
4
to
the
Daphnid,
Daphnia
Magna.
T.
R.
Wilbury
Laboratories,
Inc.,
3M
Company
Lab
Request
number
N2803­
4.
March
25.

Boeri,
R.,
Kowalski,
P.,
Ward,
T.
1996c.
Acute
Toxicity
of
N2803­
2
to
the
Daphnid,
Daphnia
magna.
T.
R.
Wilbury
Laboratories,
Inc.,
3M
Company
Lab
Request
Number
N2803­
2.
March
25.

Boyd,
S.
1993a.
Review
of
Technical
Report
Summary:
Adsorption
of
FC
95
and
FC
143
in
Soil.
Michigan
State
University.
May
19.

Boyd,
S.
A.
1993b.
Review
of
Technical
Notebook.
Soil
Thin
Layer
Chromatography.
Number
48277,
p30.
Michigan
State
University.

Bultman,
D.
and
Pike,
M.
of
3M
"
The
Use
of
Fluorochemical
Surfactants
in
Floor
Polish",
http://
home.
hanmir.
com/~
hahnw/
news/
3m.
html
Burris,
JM;
Olsen,
G;
Simpson,
C;
Mandel,
J.
2000
Determination
of
serum
half­
lives
of
several
fluorochemicals.
Interim
Report
#
1,
Corporate
Occupational
Medicine
Department,
3M
Company.

Burris,
JM,
Lundberg,
JK,
Olsen,
GW,
Simpson,
C,
Mandel,
J.
2002.
Determination
of
serum
half­
lives
of
several
fluorochemicals.
Interim
Report
#
2.
3M
Medical
Department.

Butenhoff,
J,
Costa,
G,
Elcombe,
C,
Farrar,
D,
Hansen,
K,
Iwai,
H,
Jung,
R,
Kennedy,
G,
Lieder,
P,
Olsen,
G,
and
Thomford,
P.
2002.
Toxicity
of
ammonium
perfluorooctanoate
in
male
cynomolgus
monkeys
after
oral
dosing
for
6
months.
Toxicol.
Sci.
69:
244­
257.

Cameron,
R.
G.,
Imaida,
K.,
Tsuda,
H.
and
Ito,
N.
1982.
Promotive
effects
of
steroids
and
bile
91
acids
on
hepatocarcinogenesis
initiated
by
diethylnitrosamine.
Cancer
Res.
42:
2426­
2428.

Carlfors,
J.
et.
al.
Colloid
Interface
Sci.
103,
332
­
336.(
1985)

Chandra,
M.,
Riley,
M.
G.
I.
and
Johnson,
D.
E.
1992.
Spontaneous
neoplasms
in
aged
spraguedawley
rats.
Arch.
Toxiocol.
66:
496­
502.

Chapin,
R.
E.,
Stevens,
J.
T.,
Hughes,
C.
L.,
Kelce,
W.
R.,
Hess,
R.
A.
and
Daston,
G.
P.
1996.
Symposium
overview:
endocrine
modulation
of
reproduction.
Fund.
Appl.
Toxicol.
29:
1­
17.

Chemguard,
2000.
Chemguard
Inc.,
Press
Release
Re:
Chemguard
Incorporated
Announces
Major
Breakthrough
in
AFFF
Fire
Fighting
Foam
Technology,
June
30,
2000.

Christopher,
B.
and
Marias,
A.
J.
1977.
28­
Day
Oral
Toxicity
Study
with
FC­
143
in
Albino
Mice,
Final
Report,
Industrial
Bio­
Test
Laboratories,
Inc.
Study
No.
8532­
10655,
3M
Reference
No.
T­
1742CoC,
Lot
269.

Clegg,
E.
D.,
Cook,
J.
C.
,
Chapin,
R.
E.,
Foster,
P.
M.
D.
and
Daston,
G.
P.
1997.
Leydig
cell
hyperplasia
and
adenoma
formation:
mechanisms
and
relevance
to
humans.
Reproduct.
Toxicol.
11:
107­
121.

CLOGP(
v4.71)
 
Calculation
of
hydrophobicity
as
Log
P(
o/
w).
2001.
Daylight
Chemical
Information
Systems,
Inc.
www.
daylight.
com
Cook,
J.
C.,
Hurtt,
M.
E.,
Frame,
S.
R.,
and
Biegel,
L.
B.
1994.
Mechanisms
of
extrahepatic
tumor
induction
by
peroxisome
proliferators
in
Crl:
CD
BR(
CD)
rats.
Toxicologist,
14:
301,
abstract
#
1169.

Cutts,
J.
H.
and
Noble,
R.
L.
1964.
Estrone­
induced
mammary
tumors
in
the
rat.
I.
Induction
and
behavior
of
tumors.
Cancer
Res.
24:
1116­
1123.

Daikin,
2001.
Nobuhiko
Tsuda,
Daikin
Industries
Ltd.,
"
Fluoropolymer
Emulsion
for
High­
Performance
Coatings"
in
Paint
&
Coating
Industry
Magazine,
June
2001,
p.
56­
66.

DCP,
1998.
Directory
of
World
Chemical
Producers:
1998
Edition.
Chemical
Information
Services.
Dallas,
TX.

DuPont,
2000.
Voluntary
Use
and
Exposure
Information
Profile.
DuPont
submission
to
USEPA,
dated
June
23,
2000.

Dynax,
2000.
Letter
from
Eduard
Kleiner
(
Dynax)
to
Charlie
Auer
(
USEPA),
Re:
Phase­
Out
of
3M
AFFF
Agents,
dated
August
2,
2000.

Edwards,
PJB
et.
al.
LANGMUIR
13(
10),
2665­
2669
(
1997)
92
EG&
G
Bionomics
Aquatic
Toxicology
Laboratory.
1978.
The
effects
of
continuous
exposure
to
78.03
on
hatchability
of
eggs
and
growth
and
survival
of
fry
of
fathead
minnow
(
Pimephales
promelas).
Report
#
BW­
78­
6­
175.
Research
report
submitted
to
3M
Company,
St.
Paul,
MN.

Ellis
D.
A.,
S.
A.
Mabury,
J.
W.
Martin
and
D.
C.
G.
Muir
2001.
Thermolysis
of
fluoropolymers
as
a
potential
source
of
halogenated
organic
acids
in
the
environment.
Nature:
412,
pp.
321­
324.

Elnabarawy,
M.
T.
1981.
3M
Technical
Report
Summary,
Multi­
Phase
Exposure/
Recovery
Algal
Assay
Test
Method.
Report
Number
006.
Project
Number
9970030000.
October
16.

EnviroSystems,
Inc.
1990a.
Static
Acute
Toxicity
of
FX­
1003
to
the
Fathead
Minnow,
Pimephales
promelas.
Hampton,
NH.
Study
number
was
9014­
3.

EnviroSystems,
Inc.
1990b.
Static
Acute
Toxicity
of
FX­
1003
to
the
Daphnid,
Daphnia
magna.
Hampton,
NH.
EnviroSystems
study
number
9013­
3.

ES&
T,
2000.
Cheryl
Moody
and
Jennifer
Field,
"
Perfluorinated
Surfactants
and
the
Environmental
Implications
of
Their
Use
in
Fire­
Fighting
Foams"
in
Environmental
Science
&
Technology,
Vol.
34,
Issue
18,
p.
3864­
3870.

FMG,
2001.
Verbal
comments
by
the
Fluoropolymer
Manufacturers
Group
of
the
Society
of
the
Plastics
Industry.
FMG/
EPA
meeting,
March
7,
2001.

FMG
Fluorochemical
Manufacturers
Group
AR226­
1094
2002
(
Fluoropolymer
Manufacturers
Group
Presentation
Slides).

Gabriel,
Karl.
Summary
of:
Primary
Skin
Irritation
Study
 
Rabbits.
Performed
by:
Biosearch.
Submitted
to
3M
Company,
3M
Center,
St.
Paul,
MN.

Gabriel,
Karl.
Summary
of:
Primary
Eye
Irritation
Study
 
Rabbits.
Performed
by:
Biosearch.
Submitted
to
3M
Company,
3M
Center,
St.
Paul,
MN.

Garry,
V.
F.,
and
R.
L.
Nelson.
1981.
An
Assay
of
Cell
Transformation
and
Cytotoxicity
in
C3H
10
½
Clonal
Cell
Line
for
the
Test
Chemical
T­
2942
CoC.
3M
Company,
St.
Paul,
MN.

Geisy,
J.
P.,
Dr.
1995.
3M
requested
expert
overview
of
"
Bioaccumulative
Properties
of
Ammonium
Perfluorooctanoate:
Static
Fish
Test".
Geisy
Ecotoxicology,
Inc.
March
20.

Giesy
J.
P.
and
K.
Kannan,
2001a.
Accumulation
of
perfluorooctanesulfonate
and
related
fluorochemicals
in
fish
tissues.
Prepared
for
3M,
St.
Paul
MN.
June
20.
In
U.
S.
EPA
Administrative
Record
AR226­
1030A
Giesy
J.
P.
and
K.
Kannan,
2001b.
Perfluorooctanesulfonate
and
related
fluorochemicals
in
fish­
93
eating
water
birds.
Prepared
for
3M,
St.
Paul
MN.
June
20.
In
U.
S.
EPA
Administrative
Record
AR226­
1030A
Giesy
J.
P.
and
K.
Kannan,
2001c.
Accumulation
of
perfluorooctanesulfonate
and
related
fluorochemicals
in
mink
and
river
otters.
Prepared
for
3M,
St.
Paul
MN.
June
20.
In
U.
S.
EPA
Administrative
Record
AR226­
1030A
Giesy
J.
P.
and
K.
Kannan,
2001d.
Perfluorooctanesulfonate
and
related
fluorochemicals
in
oyster,
Crassostrea
virginica,
from
the
Gulf
of
Mexico
and
Chesapeake
Bay.
Prepared
for
3M,
St.
Paul
MN.
June
20.
In
U.
S.
EPA
Administrative
Record
AR226­
1030A
Giesy
J.
P.
and
J.
L.
Newsted,
2001e.
Selected
fluorochemicals
in
the
Decatur,
Alabama
area.
Prepared
for
3M,
St.
Paul
MN,
Project
178401.
June.
In
U.
S.
EPA
Administrative
Record
AR226­
1030A
Gibson,
S.
J.,
and
Johnson,
J.
D.
1979.
Absorption
of
FC­
143­
14C
In
Rats
After
a
Single
Oral
Dose.
Riker
Laboratories,
Inc.,
Subsidiary
of
3M,
St.
Paul,
Minnesota.

Gibson,
S.
J.,
and
Johnson,
J.
D.
1980.
Extent
and
Route
of
Excretion
and
Tissue
Distribution
of
Total
Carbon­
14
in
Male
and
Female
Rats
After
a
Single
IV
Dose
of
FC­
143­
14C.
Riker
Laboratories,
Inc.,
Subsidiary
of
3M,
St.
Paul,
Minnesota.

Gibson,
S.
J.,
and
Johnson,
J.
D.
1983.
Extent
and
Route
of
Excretion
of
Total
Carbon­
14
in
Pregnant
Rats
After
a
Single
Oral
Dose
of
Ammonium
14
C­
Perfluorooctanoate.
Riker
Laboratories,
Inc.,
Subsidiary
of
3M,
St.
Paul,
Minnesota.

Gillett,
James.
1993.
3M­
requested
expert
review
of
"
Bioaccumulation
Studies".
Cornell
University.
March
8.

Gilliland,
F.
1992.
Fluorochemicals
and
Human
Health:
Studies
in
an
Occupational
Cohort.
Doctoral
thesis,
Division
of
Environmental
and
Occupational
Health,
University
of
Minnesota.

Gilliland,
F.
D.
and
Mandel,
J.
S.
1993.
Mortality
among
employees
of
a
perfluorooctanoic
acid
production
plant.
JOM.
35(
9):
950­
954.

Gilliland,
FD
and
Mandel,
JS.
1996.
Serum
Perfluorooctanoic
acid
and
hepatic
enzymes,
lipoproteins,
and
cholesterol:
a
study
of
occupationally
exposed
men.
Am
J
Ind
Med
29:
560­
568.

Glaza,
S.
1995.
Acute
dermal
toxicity
study
of
T­
6342
in
rabbits.
Corning
Hazelton,
Inc.
Madison,
WI.
Project
ID:
HWI
50800374.
3M
Company.
St.
Paul,
MN.

Glaza,
S.
M.
1997.
Acute
Oral
Toxicity
Study
of
T­
6669
in
Rats.
Corning
Hazleton
Inc.
CHW
61001760.
January
10.
Sponsored
by
3M,
St.
Paul,
Minnesota.
94
Goldenthal,
E.
I.
1978a.
Ninety
Day
Subacute
Rat
Toxicity
Study.
Final
Report
Prepared
for
3M,
St
Paul,
Minnesota,
by
International
Research
and
Development
Corporation,
St.
Paul,
Minnesota,
November
6,
1978.

Goldenthal,
E.
I.
1978b.
Ninety
Day
Subacute
Rhesus
Monkey
Toxicity
Study.
Final
Report
Prepared
for
3M,
St
Paul,
Minnesota,
by
International
Research
and
Development
Corporation,
St.
Paul,
Minnesota,
November
10,
1978.

Gortner,
E.
G.
1981.
Oral
Teratology
Study
of
T­
2998CoC
in
Rats.
Safety
Evaluation
Laboratory
and
Riker
Laboratories,
Inc.
Experiment
Number:
0681TR0110,
December
1981.

Gortner,
E.
G.
1982.
Oral
Teratology
Study
of
T­
3141CoC
in
Rabbits.
Safety
Evaluation
Laboratory
and
Riker
Laboratories,
Inc.
Experiment
Number:
0681TB0398,
February
1982.

Griffith,
F.
D.,
and
Long
J.
E.
1980.
Animal
toxicity
studies
with
ammonium
perfluorooctanoate.
Am.
Ind.
Hyg.
Assoc.
J.
41(
8):
576­
583.

Hanhijarvi,
H.,
Ophaug,
R.
H.,
and
Singer,
L.
1982.
The
sex­
related
difference
in
perfluorooctanoate
excretion
in
the
rat.
Proc.
Soc.
Exp.
Biol.
Med.
171:
50­
55.

Hanhijarvi,
H.,
M.
Ylinen,
A.
Kojo,
and
V.
Kosma.
1987.
Elimination
and
toxicity
of
perfluorooctanoic
acid
during
subchronic
administration
in
the
Wistar
rat.
Pharmacol.
Toxicol.
61:
66­
68.

Hanhijarvi,
H.
et
al.
1988.
A
proposed
species
difference
in
the
renal
excretion
of
perfluorooctanoic
acid
in
the
beagle
dog
and
rat
In:
Beynen,
A.
C.
and
H.
A.
Solleveld
(
Eds).
New
Developments
in
Biosciences:
Their
Implications
for
Laboratory
Animal
Science.
Martinus
Nijhoff
Publishers.
Dordrecht,
Netherlands.

Hansch,
C
and
Leo,
A
(
Eds.).
1979.
Chapter
IV,
The
Fragment
Method
of
Calculated
Partition
Coefficients.
Substituent
Constants
for
Correlation
Analysis
and
Chemistry
and
Biology.
John
Wiley
and
Sons,
Inc.

Hansen
K.
J.,
H.
O.
Johnson,
J.
S.
Eldridge,
J.
L.
Butenhoff,
and
L.
A.
Dick,
2002,
Quantitative
characterization
of
trace
levels
of
PFOS
and
PFOA
in
the
Tennessee
River.
Environ.
Sci.
Technol.
2002,
36,
1681­
1685.

Hatfield,
T.
2001.
Screening
Studies
on
the
Aqueous
Photolytic
Degradation
of
Perfluorooctanoic
Acid
(
PFOA).
3M
Environmental
Laboratory.
Lab
request
number
E00­
2192.
St.
Paul,
MN.

Henwood,
S.
1997.
5
Daily
Dose
Oral
Toxicity
Study
with
T­
6669
in
Rats.
Corning
Hazleton,
Inc.,
Madison,
WI.
Laboratory
Project
Identification:
CHW
6329­
197.
3M
Company,
St.
Paul,
95
MN.

Heuvel,
J.
P.
V.
et
al.
1991.
Tissue
distribution,
metabolism,
and
elimination
of
perfluorooctanoic
acid
in
male
and
female
rats.
J.
Biochem.
Toxicology.
6(
2):
83­
92.

Heuvel,
J.
P.
V.
et
al.
1992.
Renal
excretion
of
perfluorooctanoic
acid
in
male
rats:
inhibitory
effect
of
testosterone.
J.
Biochem.
Toxicology.
7(
1):
31­
36.

Heuvel,
J.
P.
V.
et
al.
1992.
Covalent
Binding
of
Perfluorinated
Fatty
Acids
to
Proteins
in
the
Plasma,
Liver,
and
Testes
of
Rats.
Chem­
Biol
Interactions.
82:
317­
328.

Howell
R.
D.,
Johnson,
J.
D.,
Drake,
J.
B.,
Youngblom,
R.
D.
1995.
Assessment
of
the
Bioaccumulative
Properties
of
Ammonium
Perfluorooctanoate:
Static.
3M
Technical
Report.
May
31.

IARC,
1995.
Peroxisome
Proliferation
and
its
Role
in
Carcinogenesis.
IARC
Technical
Report
No.
24.
International
Agency
for
Research
on
Cancer,
Lyon,
France.

Ikeda,
T.,
Aiba,
K.,
Fukuda,
K.
and
Tanaka,
M.
1985.
The
induction
of
peroxisome
proliferation
in
rat
liver
by
perfluorinated
fatty
acids,
metabolically
inert
derivatives
of
fatty
acids.
J.
Biochem.
98:
475­
482
Industrial
Bio­
Test
Laboratories,
Inc.
1977a.
Report
to
3M
Company:
28­
Day
Oral
Toxicity
Study
with
FC­
143
in
Albino
Mice.
IBT
No.
8532­
10655.

Industrial
Bio­
Test
Laboratories,
Inc.
1977b.
Report
to
3M
Company:
28­
Day
Oral
Toxicity
Study
with
FC­
143
in
Albino
Rats.
IBT
No.
8532­
10654.

Johnson,
J.
D.
1995a.
Final
Report,
Analytical
Study,
Single­
Dose
Intravenous
Pharmacokinetic
Study
of
T­
6067
in
Rabbits.
Study
Number:
AMDT­
120694.1.
3M
Environmental
Technology
&
Services,
St.
Paul,
MN.

Johnson,
J.
D.
1995b.
Final
Report,
Analytical
Study,
Single­
Dose
Absorption/
Toxicity
Study
of
T­
6067,
T­
6068,
and
T­
6069
in
Rabbits.
Study
Number:
AMDT­
011095.1.
3M
Environmental
Technology
&
Services,
St.
Paul,
MN.

Johnson,
J.
D.,
Gibson,
S.
J.,
and
Ober,
R.
E.
1984.
Cholestyramine­
enhanced
fecal
elimination
of
carbon­
14
in
rats
after
administration
of
ammonium
[
14C]
perfluorooctanoate
or
potassium
[
14C]
perfluorooctanesulfonate.
Fund.
Appl.
Toxicol.
4:
972­
976.

Kachanova,
Z.
P.;
Koslov,
J.
N.
Zh.
Fiz.
Khim.
1973,
Vol.
47,
p.
2107.

Kawashima,
Y.,
Uy­
Yu,
N.
and
Kozuka,
H.
1989.
Sex­
related
difference
in
the
inductions
by
perfluoro­
octanoic
acid
of
peroxisome
 ­
oxidation,
microsomal
1­
acylglycerolphosphocholine
96
acyltransferase
and
cytosolic
long­
chain
acyl­
CoA
hydrolase
in
rat
liver.
Biochem.
J.
261:
595­
600.

Kennedy,
G.
L.
1985.
Dermal
toxicity
of
ammonium
perfluorooctanoate.
Toxicol.
Appl.
Pharmacol.
81(
2):
348­
355.

Kennedy,
G.
L.
1987.
Increase
in
mouse
liver
weight
following
feeding
of
ammonium
perfluorooctanoate
and
related
fluorochemicals.
Toxicol.
Lett.
39:
295­
300.

Kennedy,
G.
L.,
Hall,
G.
T.,
Brittelli,
M.
R.,
Barnes,
J.
R.,
and
Chen,
H.
C.
1986.
Inhalation
toxicity
of
ammonium
perfluorooctanoate.
Food
Chem.
Toxicol.
24(
12):
1325­
1329.

Kidde,
2000.
Kidde
Fire
Fighting,
Press
Release
Re:
3M
Withdraws
from
Fire
Fighting
Foam
Manufacture,
May
30,
2000.

Kirk­
Othmer,
1994.
"
Fluorinated
Higher
Carboxylic
Acids"
under
"
Fluorine
Compounds,
Organic
(
Higher
Acids)"
in
Kirk­
Othmer
Encyclopedia
of
Chemical
Technology,
4th
ed.,
Vol.
11,
pp.
551­
558.

Kudo,
N.,
Katakura,
M.,
Sato,
Y.,
Kawashima,
Y.
2002.
Sex
hormone­
regulated
renal
transport
of
perfluorooctanoic
acid.
Chem.
Biol.
Interact.
139:
301­
316.

Kurume
Laboratory,
2001.
Bioaccumulation
test
of
Perfluoroalkylcarboxylic
acid
(
C=
7­
13)
[
This
test
is
performed
using
Perfluorooctanoic
acid
(
Test
substance
number
K­
1519)]
in
carp.
Test
No.
51519.,
Chemicals
Evaluation
and
Research
Institute,
Japan.,
December
18,
2001,
pages
1­
26.

Lawlor,
T.
E.
1996.
Mutagenicity
Test
with
T­
6564
in
the
Salmonella
 
Escherichia
Coli/
Mammalian­
microsome
Reverse
Mutation
Assay
with
a
Confirmatory
Assay.
Corning
Hazleton
Inc.
Final
Report.
CHV
Study
No:
17750­
0­
409R.
September
13.

Lawlor,
T.
1995.
Mutagenicity
test
with
T­
6342
in
the
Salmonella­
Escherichia
coli/
mammalian­
microsome
reverse
mutation
assay.
Laboratory
Number:
17073­
0­
409.
Corning
Hazleton
Inc.,
Vienna,
VA.
3M
Company.
St.
Paul,
MN.

Lines,
D.
and
Sutcliffe,
H.,
Journal
of
Fluorine
Chemistry,
25,
505
­
512
(
1984).

Liu,
R.
C.
M.,
Hurtt,
M.
E.,
Cook,
J.
C.
and
Biegel,
L.
B.
1996.
Effect
of
the
peroxisome
proliferator,
ammonium
perfluorooctanoate
(
C8),
on
hepatic
aromatase
activity
in
adult
male
Crl:
CD
BR
(
CD)
rats.
Fund.
Appl.
Toxicol.
30:
220­
228.

Liu,
S.
C.,
Sanfilippo,
B.,
Perroteau,
I.,
Derynck,
R.,
Salomon,
D.
S.
and
Kidwell,
W.
R.
1987.
Expression
of
transforming
growth
factor
 
(
TGF )
in
differentiated
rat
mammary
tumors:
estrogen
induction
of
TGF 
production.
Mol.
Endocrinol.
1:
683­
692.
97
Longnecker,
D.
S.
1987.
Interface
between
adaptive
and
neoplatic
growth
in
the
pancreas.
Gut,
28:
253­
258.

LPSD,
2000.
www.
lpsd.
zzzip.
net/
projects.
htm
Lunak,
S.;
Sediak,
P.
Photoinitiated
Reactions
of
Hydrogen
Peroxide
in
the
Liquid
Phase.
J.
Photochem.
Photobiol.
A.:
Chem.
1992,
Vol.
68,
pp.
1­
33.

Mendel,
A.
1978.
Soil
Thin
Layer
Chromatography
 
FC­
95,
FC­
143,
FM­
3422.
Excerpt
from
3M
Technical
Notebook.
October
13,
1978.
Number
48277,
p30.
Project
Number
9970612600.

Metrick,
M.
and
Marias,
A.
J.
1977.
28­
Day
Oral
Toxicity
Study
with
FC­
143
in
Albino
Rats,
Final
Report,
Industrial
Bio­
Test
Laboratories,
Inc.
Study
No.
8532­
10654,
3M
Reference
No.
T­
1742CoC,
Lot
269,
September
29,
1977.

Moody
C.
A.
and
J.
A.
Field,
1999.
Determination
of
Perfluorocarboxylates
in
Groundwater
Impacted
by
Fire­
fighting
Activity.
Environ.
Sci.
Technol.
1999,
33,
2800­
2806.

Moody,
C.
A.,
J.
W.
Martin,
W.
C.
Kwan,
D.
C.
G.
Muir,
S.
A.
Mabury.
2002.
Monitoring
Perfluorinated
Surfactants
in
Biota
and
Surface
Water
Samples
Following
an
Accidental
Release
of
Fire­
Fighting
Foam
into
Etobiocoke
Creek.
Environ.
Sci.
Technol.
2002,
36(
4),
545­
551.

Murli,
H.
1995.
Mutagenicity
test
on
T­
6342
in
an
mouse
micronucleus
assay.
Corning
Hazleton
Inc.,
Vienna,
VA.
3M
Company.
St.
Paul,
MN.

Murli,
H.
1996a.
Mutagenicity
test
on
T­
6564
in
an
mouse
micronucleus
assay.
Study
number
17750­
0­
455.
3M
Company,
St.
Paul,
MN.

Murli,
H.
1996b.
Mutagenicity
Test
on
T­
6564
Measuring
Chromosomal
Aberrations
in
Chinese
Hamster
Ovary
(
CHO)
Cells:
with
a
Confirmatory
Assay
with
Multiple
Harvests.
Final
Report.
Corning
Hazleton
Inc.
CHV
Study
No.:
17750­
0­
437CO.
September
16.

Murli,
H.
1996c.
Mutagenicity
Test
on
T­
6342,
Measuring
Chromosomal
Aberrations
in
Whole
Blood
Lymphocytes
With
a
Confirmatory
Assay
With
Multiple
Harvests.
Corning­
Hazelton,
Inc.
(
CHV).
Vienna,
VA.
CHV
Study
No.:
17073­
0­
449CO.

Murli,
H.
1996d.
Mutagenicity
Test
on
T­
6342,
Measuring
Chromosomal
Aberrations
in
Chinese
Hamster
Ovary
(
CHO)
Cells:
with
a
Confirmatory
Assay
with
Multiple
Harvests.
Corning­
Hazelton,
Inc.
(
CHV).
Vienna,
VA.
CHV
Study
No.:
17073­
0­
437CO.

Napoli
M.,
Fraccaro
C.,
Scipioni
A.,
and
Armelli
R.
1984.
Thermal
Decomposition
of
Perfluoroalkanesulfonyl
Fluorides:
The
Pyrolysis
of
Perfluoro­
n­
octane­
1­
sulfonyl
Fluoride.
Journal
of
Fluorine
Chemistry,
24
,
377­
385.
98
Nilsson,
R.,
Beije,
B.,
Preat,
V.,
Erixon,
K.
and
Ramel,
C.
1991.
On
the
mechanism
of
the
hepatocarcinogenicity
of
peroxisome
proliferators.
Chem.­
Biol.
Interact.
78:
235­
250.

NOTOX.
2000.
Evaluation
of
the
Ability
of
T­
7524
to
Induce
Chromosome
Aberrations
in
Cultured
Peripheral
Human
Lymphocytes.
NOTOX
Project
Number
292062.
Hertogenbosch,
The
Netherlands.

NOTOX.
2001.
Assessment
of
Contact
Hypersensitivity
to
T­
7524
in
the
Albino
Guinea
Pig
(
Maximisation­
Test).
NOTOX
Project
number
292027.
Hertogenbosch,
The
Netherlands.

Nubbe,
M.
E.;
Adams,
V.
D.;
Moore,
W.
M.
The
Direct
and
Sensitized
Photo­
oxidation
of
Hexachlorocyclopentadiene,
Wat.
Res.
1995,
Vol.
29,
No.
5,
pp1287­
1293.

Obourn,
J.
D.,
Frame.
S.
R.,
Bell,
R.
H.
Jr.,
Longnecker,
D.
S.,
Elliott,
G.
S.
and
Cook,
J.
C.
1997.
Mechanisms
for
the
pancreatic
oncogenic
effects
of
the
peroxisome
proliferator
Wyeth­
14,643.
Toxicol.
Appl.
Pharmacol.
145:
425­
436.

OECD
Guideline
for
Testing
of
Chemicals,
Phototransformation
of
Chemicals
in
Water­
Direct
and
Indirect
Photolysis,
(
Draft
Document);
OECD,
2000,
pp1­
59.

Ogata,
Y.;
Tomizawa,
K.;
Furuta,
K.
Chemistry
of
Peroxides,
in
S.
Patai
(
ed),
The
Chemistry
of
Peroxides
1983,
p.
720
Olsen,
G.
W.
et
al.
1998a.
An
Epidemiologic
Investigation
of
Reproductive
Hormones
in
Men
with
Occupational
Exposure
to
Perfluorooctanoic
Acid.
JOEM.
40(
7):
614­
622.

Olsen,
G.
W.,
et
al.
1998b.
3M
Final
Report:
An
epidemiologic
investigation
of
plasma
cholecystokinin,
hepatic
function
and
serum
perfluorooctanoic
acid
levels
in
production
workers.
3M
Company,
St.
Paul.
Sept
4.

Olsen,
G.
W.,
Burris,
J.
M.,
Burlew,
M.
M.,
Mandel,
J.
H.
2000.
Plasma
cholecystokinin
and
hepatic
enzymes,
cholesterol
and
lipoproteins
in
ammonium
perfluorooctanoate
production
workers.
Drug
Chem
Tox.
23(
4):
603­
620.

Olsen,
GW,
Logan,
PW,
Simpson,
CA,
Burris,
JM,
Burlew,
MM,
Lundberg,
JK,
Mandel,
JH.
2001a.
Descriptive
summary
of
serum
fluorochemical
levels
among
employee
participants
of
the
year
2000
Decatur
fluorochemical
medical
surveillance
program.
Final
Report.
March
19,
2001.

Olsen,
GW,
Schmickler,
M,
Tierens,
JM,
Logan,
PW,
Burris,
JM,
Burlew,
MM,
Lundberg,
JK,
Mandel,
JH.
2001b.
Descriptive
summary
of
serum
fluorochemical
levels
among
employee
participants
of
the
year
2000
Antwerp
fluorochemical
medical
surveillance
program.
Final
Report.
March
19,
2001.
99
Olsen,
GW,
Madsen,
DC,
Burris,
JM,
Mandel,
JH.
2001c.
Descriptive
summary
of
serum
fluorochemical
levels
among
236
building
employees.
Final
Report.
March
19,
2001.

Olsen,
GW,
Hansen,
Clemen,
LA,
Burris,
JM,
Mandel,
JH.
2001d.
Identification
of
Fluorochemicals
in
Human
Tissue.
Final
Report.
Epidemiology,
220­
3W­
05,
Medical
Department,
3M
Company,
St.
Paul,
MN
55144.

Olsen
GW,
Burlew
MM,
Burris
JM,
Mandel
JH.
2001e.
A
cross­
sectional
analysis
of
serum
perfluorooctanesulfonate
(
PFOS)
and
perfluorooctanoate
(
PFOA)
in
relation
to
clinical
chemistry,
thyroid
hormone,
hematology
and
urinalysis
results
from
male
and
female
employee
participants
of
the
2000
Antwerp
and
Decatur
fluorochemical
medical
surveillance
program.
Final
report.
3M
Medical
Department.

Olsen,
G.
W.,
Burlew,
M.
M,
Burris,
J.
M.,
Mandel,
J.
H.
2001f.
A
Longitudinal
Analysis
of
Serum
Perfluorooctanesulfonate
(
PFOS)
and
Perfluorooctanoate
(
PFOA)
Levels
in
Relation
to
Lipid
and
Hepatic
Clinical
Chemistry
Test
Results
from
Male
Employee
Participants
of
the
1994/
95,
1997,
and
2000
Fluorochemical
Medical
Surveillance
Program.
3M
Final
Report.

Olsen,
GW,
Burlew,
MM,
Hocking,
BB,
Skratt,
JC,
Burris,
JM,
Mandel,
JH.
2001g.
An
epidemiologic
analysis
of
episodes
of
care
of
3M
Decatur
chemical
and
film
plant
employees,
1993­
1998.
Final
Report.
May
18,
2001.

Olsen,
GW,
Burris,
JM,
Lundberg,
JK,
Hansen,
KJ,
Mandel,
JH,
Zobel,
LR.
2002a.
Identification
of
fluorochemicals
in
human
sera.
I.
American
Red
Cross
adult
blood
donors.
Final
report.
3M
Medical
Department.

Olsen,
GW,
Burris,
JM,
Lundberg,
JK,
Hansen,
KJ,
Mandel,
JH,
Zobel,
LR.
2002b.
Identification
of
fluorochemicals
in
human
sera.
II.
Elderly
participants
of
the
Adult
Changes
in
Thought
Study,
Seattle,
Washington.
Final
Report.
3M
Medical
Department.

Olsen,
GW,
Burris,
JM,
Lundberg,
JK,
Hansen,
KJ,
Mandel,
JH,
Zobel,
LR.
2002c.
Identification
of
fluorochemicals
in
human
sera.
III.
Pediatric
participants
in
a
Group
A
Streptococci
clinical
trial
investigation.
Final
Report.
3M
Medical
Department.

O'Malley,
K.
D.,
and
Ebbens,
K.
L.
1981.
Repeat
Application
28
Day
Percutaneous
Absorption
Study
with
T­
2618CoC
in
Albino
Rabbits.
Riker
Laboratories,
St.
Paul,
MN.

Ophaug,
R.
H.
and
L.
Singer.
1980.
Metabolic
Handling
of
Perfluorooctanoic
Acid
in
Rats.
Proc
Soc
Exp
Biol
Med.
163:
19­
23.

Pace
Analytical.
1997.
Ready
Biodegradation
of
FC­
126(
BOD/
COD).
3M
Company
Lab
Request
No.
E1282.
Minneapolis,
MN.
May
29.
100
Pace
Analytical.
2001.
The
18­
Day
Aerobic
Biodegradation
Study
of
Perfluorooctanesulfonyl­
Based
Chemistries.
3M
Company
Request,
Contract
Analytical
Project
ID:
CA097,
Minneapolis,
MN.
February
23.

Palazzolo,
M.
J.
1993.
Thirteen­
Week
Dietary
Toxicity
Study
with
T­
5180,
Ammonium
Perfluorooctanoate
(
CAS
No.
3825­
26­
1)
in
Male
Rats.
Final
Report.
Laboratory
Project
Identification
HWI
6329­
100.
Hazleton
Wisconsin,
Inc.

Pastoor,
T.
P.,
Lee,
K.
P.,
Perri,
M.
A.
and
Gillies,
P.
J.
1987.
Biochemical
and
morphological
studies
of
ammonium
perfluorooctanoate­
induced
hepatomegaly
and
peroxisome
proliferation.
Exp.
Mol.
Pathol.
47:
98­
109.

Petritis,
1999.
K.
Petritis,
et
al.
"
Ion­
pair
reversed­
phase
liquid
chromatography
for
determination
of
polar
underivatized
amino
acids
using
perfluorinated
carboxylic
acids
as
ion
pairing
agent"
in
J.
Chromatography
A,
Vol.
833,
1999,
pp.
147­
155.

Reiner,
E.
A.
1978.
Fate
of
Fluorochemicals
in
the
Environment.
Project
Number
9970612613.
3M
Company,
Environmental
Laboratory.
July
19.

Reiner,
E.
A.
1981.
3M
Company
Environmental
Laboratory,
St.
Paul,
Minnesota,
Dec.
7.

Renner,
2001.
"
Growing
Concern
Over
Perfluorinated
Chemicals"
in
Environmental
Science
and
Technology,
Vol.
35,
Issue
7,
pp.
154A­
160A,
April
1,
2001.

Riker
Laboratories,
Inc.,
Safety
Evaluation
Laboratory.
1979.
Repeat
Application
28­
Day
Percutaneous
Absorption
Study
with
T­
2618CoC
in
Albino
Rabbits.
St.
Paul,
Minnesota.
Experiment
Number:
09790AB0485.

Riker
Laboratories
Inc.,
Safety
Evaluation
Laboratory.
1983.
Primary
Skin
Irritation
Test
with
T­
3371
in
Albino
Rabbits.
St.
Paul,
MN.
Experiment
#
0883EB0079.

Sadhu,
D.
2002
CHO/
HGPRT
forward
mutation
assay
 
ISO
(
T6.889.7)
Toxicon
Corporation,
Bedford,
MA,
Toxicon
Final
Report:
01­
7019­
G1
Submitted
to:
3M,
St.
Paul
Minnesota
55144­
1000.

Scrano,
L.;
Bufo,
S.
A.;
Perucci,
P.;
Meallier,
P.;
Mansour,
M.
Photolysis
and
Hydrolysis
of
Rimsulfuron,
Pestic.
Sci.
1999,
Vol.
55,
pp.
955­
961.

Simister,
E.
et.
al.
J.
Chem.
Soc.,
Faraday
Trans.
88(
20),
3033­
41
(
1992)

Sohlenius,
A.
K.,
Anderson,
K.
and
DePierre,
J.
1992.
The
effects
of
perfluorooctanoic
acid
on
hepatic
peroxisome
proliferation
and
related
parameters
show
no
sex­
related
differences
in
mice.
Biochem.
J.
265:
779­
783.
101
Staples,
R.
E.,
Burgess,
B.
A.,
and
Kerns,
W.
D.
1984.
The
Embryo­
Fetal
Toxicity
and
Teratogenic
Potential
of
Ammonium
Perfluorooctanoate
(
APFO)
in
the
Rat.
Fund.
Appl.
Tox.
4,
429­
440.

Takagi,
A.,
Sai,
K.,
Ummemura,
T.,
Hasegawa,
R.
and
Kurokawa,
Y.
1991.
Short­
term
exposure
to
the
peroxisome
proliferators,
perfluorooctanoic
acid
and
perfluorodecanoic
acid,
causes
significant
increases
of
8­
hydroxydeoxyguanosine
in
liver
DNA
of
rats.
Cancer
Lett.
57:
55­
60.

Takagi,
A.,
Sai,
K.,
Ummemura,
T.,
Hasegawa,
R.
and
Kurokawa,
Y.
1992.
Hepatomegaly
is
an
early
biomarker
for
hepatocarcinogenesis
induced
by
peroxisome
proliferators.
J.
Environ.
Toxicol.
Pathol.
11:
145­
149.

Taylor
P.,
Dellinger
B.
and
Lee
C.
C.,
1990.
Development
of
a
Thermal
Stability
Based
Ranking
of
Hazardous
Organic
Compound
 
Incinerability.
Environmental
Science
&
Technology,
24,
316­
328.

Teerds,
K.
J.,
Rommerts,
F.
G.
and
Dorrington,
J.
H.
1990.
Immunohistochemical
detection
of
transforming
growth
factor­
 
in
Leydig
cell
during
the
development
of
the
rat
testis.
Molec.
Cell
Endocrinol.
69:
R1­
R6.

Thomford,
PJ.
2001b.
26­
Week
Capsule
Toxicity
Study
with
Ammonium
Perfluorooctanoate
(
APFO)
in
Cynomolgus
Monkeys.
Study
performed
by
Covance
Laboratories
Inc.,
Madison
Wisconsin
53704­
2592
for
APME
Ad­
hoc
APFO
Toxicology
Working
Group.
Study
No.
Covance
6329­
231,
Completion
Date
December
18,
2001,
463
pp.

Thomford
PJ.
2001b.
4­
Week
Capsule
Toxicity
Study
with
Ammonium
Perfluorooctanoate
(
APFO)
in
Cynomolgus
Monkeys.
Study
performed
by
Covance
Laboratories
Inc.,
Madison
Wisconsin
53704­
2592
for
APME
Ad­
hoc
APFO
Toxicology
Working
Group.
Study
No.
Covance
6329­
230,
Completion
Date
December
18,
2001,
159
pp.

Todd,
J.
W.
1979.
FC­
143
Photolysis
Study
Using
Simulated
Sunlight.
Project
9776750202,
3M
Company
Technical
Report
No.
002.
February
2.

T.
R.
Wilbury
Laboratories,
Inc.
1995.
Growth
and
Reproduction
Toxicity
Test
with
N2803­
3
and
the
Freshwater
Alga,
Selenastrum
capricornutum.
Marblehead,
MA.
Study
number
B93­
TH.

T.
R.
Wilbury
Laboratories,
Inc.
1996a.
Acute
toxicity
of
N2803­
3
to
the
Fathead
Minnow,
Pimephales
promelas.
Marblehead,
MA.
Study
number
891­
TH.

T.
R.
Wilbury
Laboratories,
Inc.
1996b.
Acute
Toxicity
of
N2803­
3
to
the
Daphnid,
Daphnia
magna.
Marblehead,
MA.
Study
number
892­
TH.

Tsang
W.,
Burgess
D.
R.,
and
Babushok
V.,
1998.
On
the
Incinerability
of
Highly
Fluorinated
Organic
Compounds.
Combustion
Science
&
Technology,
1998,
139,
385­
402.
102
Ubel
FA,
Sorenson
SD,
Roach
DE.
1980
Health
status
of
plant
workers
exposed
to
fluorochemicals­­
a
preliminary
report.
Am
Ind
Hyg
Assoc
J
41:
584­
589.

USEPA
1998
Fate,
Transport
and
Transformation
Test
Guidelines,
OPPTS
835.5270
Indirect
Photolysis
Screening
Test;
EPA712­
C­
98­
099;
United
States
Environmental
Protection
Agency,
U.
S.
Government
Printing
Office:
Washington,
DC,
1998,
pp1­
22.

USEPA
1998.
Fate,
Transport
and
Transformation
Test
Guidelines:
835.2110:
Hydrolysis
as
a
Function
of
pH;
EPA712­
C­
98­
057;
United
States
Environmental
Protection
Agency,
U.
S.
Government
Printing
Office:
Washington,
DC.

USEPA
2002a.
Memorandum
from
Dr.
Ralph
Cooper,
NHEERL,
to
Dr.
Jennifer
Seed,
dated
October
2,
2002.

USEPA
2002b.
Memorandum
from
Dr.
Elizabeth
Margosches
to
Dr.
Katherine
Anitole,
dated
October
21,
2002.

Vanden
Heuvel,
J.
P.,
Davis,
J.
W.,
Sommers,
R.,
and
Peterson,
R.
E.
1992.
Renal
excretion
of
perfluorooctanoic
acid
in
male
rats:
Inhibitory
effect
of
testosterone.
J.
Biochem.
Toxicol.
7(
1):
31­
36.

Vanden
Heuvel,
J.
P.,
Kuslikis,
B.
I.,
and
Peterson,
R.
E.
1991a.
Covalent
binding
of
perfluorinated
fatty
acids
to
proteins
in
the
plasma,
liver
and
testes
of
rats.
Chem.­
Biol.
Interact.
82:
317­
328.

Vanden
Heuvel,
J.
P.,
Kuslikis,
B.
I.,
Van
Rafelghem,
M.
L.
and
Peterson,
R.
E.
1991b.
Tissue
distribution,
metabolism,
and
elimination
of
perfluorooctanoic
acid
in
male
and
female
rats.
J.
Biochem.
Toxicol.
6(
2):
83­
92.

Vraspir,
G.
A.,
Mendel,
Arthur.
1979.
Analysis
for
Fluorochemicals
in
Bluegill
Fish.
Project
9970612600:
Fate
of
Fluorochemicals.
3M
Technical
Report
Report
Number
14.
May
1.

Ward,
T.,
Nevius,
J.
and
R.
Boeri.
1996a.
Acute
toxicity
of
FC­
1015
to
the
fathead
minnow,
Pimephales
promelas.
T.
R.
Wilbury
Laboratories,
Inc.
Lab
Request
number
P1624.
3M
Company,
St.
Paul,
MN.

Ward,
T.,
Nevius,
J.
and
R.
Boeri.
1996b.
Growth
and
reproduction
toxicity
test
with
FC­
1015
and
the
freshwater
alga,
Selenastrum
capricornutum.
T.
R.
Wilbury
Laboratories,
Inc.
Lab
Request
number
P1624.
3M
Company,
St.
Paul,
MN.

Ward,
T.,
Nevius,
J.,
and
R.
Boeri.
1996c.
Acute
toxicity
of
FC­
1015
to
Daphnid,
Daphnia
magna.
T.
R.
Wilbury
Laboratories,
Inc.
Lab
request
number
P1624.
3M
Company,
St.
Paul,
MN.
103
Welsh,
S.
K.
1978.
Technical
Report
Summary
­
Adsorption
of
FC­
95
and
FC­
143
on
soil.
Environmental
Laboratory.
3M
Company
Project
9970612633:
Fate
of
Fluorochemicals,
Report
Number
1.
St.
Paul,
MN.
February
27.

Yager,
J.
D.
Jr.
and
Yager,
R.
1980.
Oral
contraceptive
steroids
as
promoters
of
hepatocarcinogenesis
in
female
Sprague­
Dawley
rats.
Cancer
Res.
40:
3680­
3685.

Yang,
Q.,
Xie,
Y.,
and
Depierre,
W.
2000.
Effects
of
peroxisome
proliferators
in
the
thymus
of
spleen
and
mice.
Clin.
Exp.
Immunol.
122:
219­
226.

Yang,
Q.,
Xie,
Y.,
Ericksson,
A.
M.,
Nelson,
B.
D.,
and
DePierre,
J.
W.
2001.
Further
evidence
for
the
involvement
of
inhibition
of
cell
proliferation
and
development
in
thymic
and
splenic
atrophy
induced
by
the
peroxisome
proliferator
perfluoroctanoic
acid
in
mice.
Biochem.
Pharmacol.
62:
1133­
1140.

Yang,
Q.,
Xie,
Y.,
Alexson,
S.
E.
H.,
Nelson,
B.
D.,
and
DePierre,
J.
W.
2002a.
Involvement
of
the
peroxisome
proliferator­
activated
receptor
alpha
in
the
immunomodulation
caused
by
peroxisome
proliferators
in
mice.
Biochem.
Pharmacol.
63:
1893­
1900.

Yang,
Q.,
Abedi­
Valugerdi,
M.,
Xie,
Y.
Zhao,
X.,
Moller,
G.,
Nelson,
B.
D.,
and
DePierre,
J.
W.
2002b.
Potent
suppression
of
the
adaptive
immune
response
in
mice
upon
dietary
exposure
to
the
potent
peroxisome
proliferator,
perfluorooctanoic
acid.
International
Immunopharmacology
2,
389­
397.

Ylinen,
M.,
Hanhijarvi,
H.,
Jaakonaho,
I.,
and
Peura,
P.
1989.
Stimulation
by
estradiol
of
the
urinary
excretion
of
perfluorooctanoic
acid
in
the
male
rat.
Pharmacol.
Toxicol.
65:
274­
277.

Ylinen,
M.,
Kojo,
A.,
Hanhijdrvi,
H.
and
Peura,
P.
1990.
Disposition
of
perfluorooctanoic
acid
in
the
rat
after
single
and
subchronic
administration.
Bull.
Environ.
Contam.
Toxicol.
44:
46­
53.

York,
R.
G.
2002.
Oral
(
Gavage)
Two­
Generation
(
One
Litter
Per
Generation)
Reproduction
Study
of
Ammonium
Perfluorooctanoic
(
APFO)
in
Rats.
Argus
Research
Laboratories,
Inc.
Protocol
Number:
418­
020,
Sponsor
Study
Number:
T­
6889.6,
March
26,
2002
