Document ID: EPA-HQ-OPPT-2002-0051-0004
Agency: epa
Document Type: Supporting & Related Material
Title: 
Posted Date: 2006-03-07T05:00Z

UNITED
STATES
ENVIRONMENTAL
PROTECTION
AGENCY
4422
6
­
10
93
WASHINGTON,
D.
C.
20460
OFFICE
OF
PREVENTION.
PESTICIDES
AND
TOXIC
SUBSTANCES
Dear
Interested
Party:

On
March
28,
2002,
EPA
made
available
copies
in
PDF
tile
format
(389KB)
of
the
DraJi
H~~
zard
A.
sse.
ssment
Qf'Perjluorooctunoic
Acid
(PFOA)
And
its
S&
s,
prepared
by
the
Risk
Assessment
Division
of
the
EPA
Office
of
Pollution
Prevention
and
Toxics.
EPA
has
become
aware
that
this
document
included
an
error
in
the
text
appearing
on
page
8
in
section
2.0
of
the
document,
"Production
of
PFOA
and
its
Salts."

This
error
has
now
been
rectified,
and
a
corrected
copy
of
the
full
document
is
attached.
Please
replace
your
earlier
version
of
the
draft
assessment
with
this
one,
or
print
only
the
cover
page,
the
Table
of
Contents,
and
page
8
from
this
corrected
file,
and
use
them
to
replace
those
pages
from
the
original
tile
printout.
An
errata
sheet
showing
the
page
8
deletions
in
stritnaat
and
new
revised
text
in
bold
is
also
attached
for
your
convenience
so
that
you
can
see
immediately
where
the
corrections
were
made.

We
regret
any
confusion
or
inconvenience
this
may
have
caused.

If
you
have
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questions
or
comments
concerning
the
Assessment,
please
contact
Jennifer
Seed
by
phone
at
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564­
7634,
or
by
email
at
seed.
iennifer@
epa.
gov.
If
you
wish
to
receive
a
copy
of
the
Annex
to
the
Assessment,
which
contains
robust
summaries
of
the
studies
reviewed
in
the
Assessment,
or
if
you
have
any
difficulties
opening
these
tiles,
please
contact
Mary
Dominiak
by
phone
at
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564­
8104,
by
fax
at
202­
564­
4775,
or
by
email
at
dominiak.
marv@
eua.
eov.

Sincerely,

IS/

Charles
M.
Auer,
Director
Chemical
Control
Division
Attachments
CONTAIN
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CR!

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I
ERRATA
SHEET
Corrected
411512002
Table
1.
Reported
Physicochemical
Properties
Compound
CAS
REG
#
MP
BP
VP
Sol.­
HZ0
Log
P
Rf­
C(=
O)
F
335­
64­
8
131
c
Rf­
C02H
335­
67­
l
55
c
189C
10mmHg
3.4
g/
L
Rf­
C02­
20
g/
L
NH4+
3825­
26­
l
130
c
sublimes
1
x
lOE­
5
gels
<5
RfC
O)
OMe
376­
27­
2
159
c
pH
(1
g
free
acid
/L
Water)
=
2.6
Free
acid
pKa
is
approximately
0.6
Sodium
or
Silver
salts
of
PFOA
decompose
above
250
C
to
generate
perfluoroolefins.

2.0
Production
of
PFOA
and
its
Salts
PFOA
is
commerciallv
manufactured
by
two
major
alternative
processes:
1)
the
Simons
Electra­
1
Chemical
Fluorination
(ECF)
process
or
2)
the
telomerization
process.

In
the
ECF
process,
an
electric
current
is
passed
through
a
solution
of
anhydrous
hydrogen
.
fluoride
and
an
organic
feedstock
of
octanoic
acid
or
a
derivativv.
I
The
ECF
process
replaces
the
carbon­
hydrogen
bonds
on
molecules
of
the
organic
feedstock
with
carbon­
fluorine
bonds,
in
an
identical
manner
used
to
make
PFOS.
Perfluorination
I
occurs
when
all
the
carbon­
hydrogen
bonds
are
replaced
with
carbon­
fluorine
bonds.
The
ECF
process
yields
between
30­
45
percent
straight
chain
(normal)
perfluorooctanonvl
fluoride
.
(PFOF))
along
with
a
variable
mixture
of
byproducts
and
impurities.
The
output
of
the
ECF
process
is
no;
a
pure
chemical,
but
instead
a
mixture
of
isomers
and
homologues
including
higher
and
lower
straight­
chain
homologues;
branched­
chain
perfluoroalkyl
fluorides
of
various
chain
lengths;
straight­
chain,
branched,
and
cyclic
perfluroalkanes
and
ethers;
and
other
byproducts
(3M
Company,
2000a).
After
disposal
or
recovery
of
some
of
the
byproducts
and
impurities,
the
acid
fluoridePBSF
is
base
hydrolyzed
in
1
batch
reactors
to
yield
PFOA.
The
PFOA
salts
are
synthesized
by
base
neutralization
of
the
acid
to
the
salt
in
a
separate
reactor
(3M
Company,
2000b).

In
the
telomerization
process,
tetrafluoroethylene
is
reacted
with
other
fluorine­
bearing
chemicals
to
yield
fluorinated
carboxylic
acids.
This
process
yields
pure
straight­
chain
acids
with
an
even
number
of
carbon
atoms.
Distillation
can
be
used
to
obtain
pure
components
(ECT,
1994).
Commercial
products
manufactured
through
the
telomerization
process
are
generally
mixtures
of
perfluorinated
compounds
with
even
carbon
numbers
(Renner,
200
1).

3M
Company
is
the
largest
manufacturer
and
importer
of
PFOA
and
its
salts
in
the
United
States,
3M
has
characterized
its
manufacture
of
PFOA
and
its
ammonium
and
sodium
salts
in
1997
at
less
than
500,000
kg
per
year,
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
8
CONTAIN
NO
Cl3
DRAFT
HAZARD
ASSESSMENT
OF
PERFLUOROOCTANOIC
ACID
AND
ITS
SALTS
U.
S.
Environmental
Protection
Agency
Office
of
Pollution
Prevention
and
Toxics
Risk
Assessment
Division
February
20,2002
(Corrected
April
15,2002)

3
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
July
200
1,

A
two­
generation
reproductive
toxicity
study
of
APFO
is
currently
being
conducted
and
will
be
available
in
the
spring
of
2002.
Effects
were
observed
in
a
two­
generation
reproductive
toxicity
study
of
a
related
compound,
perfluorooctane
sulfonate.
The
results
of
the
APFO
study
will
be
important
to
determine
whether
similar
effects
are
observed.
Corrected
4/
l
512002
Table
of
Contents
Executive
Summary
1.0
Chemical
Identity
1.1
Physicochemical
Properties
2.0
Production
of
PFOA
and
its
Salts
(Corrected
April
15,2002)
2.1
Uses
of
PFOA
and
its
Salts
2.2
Environmental
Fate
2.2.1
Photolysis
2.2.2
Volatility
2.2.3
Biodegradation
2.2.4
Hydrolysis
2.2.5
Bioaccumulation
2.2.6
Soil
Adsorption
2.3
Environmental
Exposure
2.3.1
Combustion
2.3.2
Discharge
to
Water
2.3.3
Discharge
to
Land
2.3.4
Environmental
Monitoring
2.4
Human
Biomonitoring
3.0
Human
Health
Hazards
3.1.
Metabolism
and
Pharmacokinetics
3.1.1
Half­
life
in
Humans
3.1.2
Absorption
Studies
in
Animals
3.1.3
Distribution
Studies
in
Animals
3.1.4
Metabolism
Studies
in
Animals
3.15
Elimination
Studies
in
Animals
3.2
Epidemiology
Studies
3.2.1
Mortality
Study
3.2.2
Hormone
Study
3.2.3
Cholesterol
Study
3.2.4
Study
on
Episodes
of
Care
(Morbidity)
3.3
Acute
Toxicity
Studies
in
Animals
3.3.1
Oral
Studies
3.3.2
Inhalation
Studies
3.3.3
Dermal
Studies
3.3.4
Eye
Irritation
Studies
3.3.5
Skin
Irritation
Studies
3.4
Mutagenicity
Studies
3.5
Subchronic
Toxicity
Studies
in
Animals
3.6
Developmental
Toxicity
Studies
in
Animals
3.7
Carcinogenicity
Studies
in
Animals
1
6
6
8
10
11
11
11
12
12
13
14
14
14
14
15
15
16
20
20
20
21
22
25
26
29
29
32
35
36
38
38
38
39
39
39
39
40
48
52
5
3.7.1
Cancer
Bioassays
52
3.7.2
Mode
of
Action
Studies
53
3.7.2.1
Liver
Tumors
53
3.7.2.2
Leydig
Cell
Tumors
54
3.7.2.3
Mammary
Gland
Tumors
55
3.7.2.4
Pancreatic
Tumors
55
4.0
Hazards
to
the
Environment
55
4.1
Introduction
55
4.2
Acute
Toxicity
to
Freshwater
Species
57
5.0
References
62
ANNEX
I
­
Robust
Summaries
77
Introduction
EXECUTIVE
SUMMARY
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
has
very
low
volatility
and
vapor
pressure.
It
does
not
hydrolyze,
photolyze
or
biodegrade
under
environmental
conditions.

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
at
a
concentration
range
of
~18
to
108
rig/
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
rig/
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
two
locations
upstream
and
five
locations
downstream
of
the
3M
manufacturing
facility
at
Decatur
AL
have
been
determined.
Of
the
five
downstream
sampling
locations,
the
two
closest
to
the
facility
had
PFOA
surface
water
concentrations
significantly
greater
than
the
two
upstream
sites
(means
of
19OOug/
L
and
1024
ug/
L);
the
nearest
three
locations
had
sediment
concentrations
significantly
greater
than
the
upstream
sites
(wet
wt.
means
1855
ug/
kg,
892
@kg,
238
ug/
kg).
The
average
fish
whole
body
PFOA
concentration
for
the
two
upstream
locations
was
11.7
ugikg
(wet
wt.),
while
that
for
the
five
downstream
locations
was
106.4
ug/
kg.
The
average
PFOA
concentration
in
clams
at
the
two
upstream
locations
was
4.38
ug/
kg,
while
the
average
for
the
five
downstream
locations
was
8.42
ug/
kg.

Based
on
available
data,
APFO
does
not
appear
to
bioaccumulate
in
fish.
In
a
study
of
fathead
minnows,
the
calculated
BCF
for
APFO
was
1.8.

7
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
85Oh)
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
LC5Os
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
>lOOO
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
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.

Human
Health
Effects
and
Biomonitoring
Little
information
is
available
concerning
the
pharmacokinetics
of
APFO
in
humans.
A
preliminary
study
of
retired
workers
suggests
simply
that
the
serum
half­
life
is
between
1
and
3.5
years.
These
data
provide
evidence
of
the
potential
to
bioaccumulate
PFOA
in
humans.
In
addition,
this
study
provides
preliminary
evidence
that
the
serum
half­
life
may
be
longer
in
females
than
in
males.

Animal
studies
have
shown
that
APFO
is
well
absorbed
following
oral
and
inhalation
exposure,
and
to
a
lesser
extent
following
dermal
exposure.
In
rats
and
dogs,
there
are
major
gender
differences
in
the
distribution
and
elimination
of
APFO.
APFO
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.
APFO
binds
to
macromolecules
in
the
tissues
listed
above.
APFO
is
not
metabolized
and
there
is
evidence
of
enterohepatic
circulation
of
the
compound.
The
urine
is
the
major
route
of
excretion
of
APFO
in
the
female
rat,
while
the
urine
and
feces
are
both
major
routes
of
excretion
of
APFO
in
male
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.
In
beagle
dogs,
the
plasma
half­
life
is
254
h
in
females
and
507
h
in
males.
In
rats,
the
elimination
half­
life
is
one
day
in
females
and
15
days
in
males.
Female
rats
appear
to
have
a
secretory
mechanism
that
rapidly
eliminates
APFO;
this
secretory
mechanism
is
either
lacking
or
relatively
inactive
in
males.
Other
studies
in
rats
have
shown
that
testosterone
exerts
an
inhibitory
effect
on
renal
excretion
of
APFO.
Hormonal
changes
during
pregnancy
do
not
appear
to
change
the
rate
of
elimination
in
rats.
The
gender
difference
observed
in
rats
and
dogs
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
mean
serum
levels
in
workers
range
from
0.84
to
6.4
ppm.
The
highest
level
reported
in
a
worker
in
1997
was
8
1.3
ppm.
In
non­
occupational
populations,
serum
PFOA
levels
were
much
lower.
In
both
pooled
blood
bank
samples
and
in
individual
samples
in
both
adults
and
children,
mean
PFOA
levels
ranged
from
3
to
17
ppb.
The
highest
serum
PFOA
level
reported
was
in
a
sample
from
a
child
(56
ppb).

Epidemiological
studies
on
the
effects
of
PFOA
in
humans
have
been
conducted
on
workers.
Two
mortality
studies,
as
well
as
studies
examining
effects
on
the
liver,
pancreas,
endocrine
system,
and
lipid
metabolism,
have
been
conducted
to
date.
In
addition,
a
morbidity
study
was
also
recently
submitted.

A
retrospective
cohort
mortality
study
demonstrated
a
weak
association
with
PFOA
exposure
and
prostate
cancer.
A
statistically
significant
association
was
observed
in
prostate
cancer
mortality
as
length
of
employment
increased.
This
result
was
not
observed
in
a
recent
update
to
the
study;
however,
the
results
cannot
be
directly
compared
because
the
exposure
categories
were
modified
in
the
update.
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
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,
except
among
obese
workers
(aspartate
amino
transferase
and
alanine
amino
transferase).
However,
PFOA
was
not
measured
directly,
but
indirectly
as
total
serum
fluorine.

3
9
There
are
many
limitations
to
these
studies,
but
most
notably
the
small
number
of
workers
with
PFOA
serum
levels
greater
than
10
ppm.
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
<lo00
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
chromosomal
aberrations
in
vitro
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
vitro.
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
CjH
lOTt/
mouse
embryo
fibroblasts
and
in
the
in
vivo
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
and
that
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/
lg/
day).
Analyses
of
serum
and
liver
levels
of
APFO
showed
a
marked
gender
difference
that
accounts
for
the
difference
in
sensitivity.
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
serum
and
liver
levels
did
not
reveal
a
gender
difference
in
monkeys,
but
the
sample
size
was
very
small
(N=
2).
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.

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
4
to
50
mg/
kg/
day
APFO.
There
was
no
evidence
of
maternal
toxicity
at
50
mg/
kg/
day,
the
highest
dose
tested.

A
two­
generation
reproductive
toxicity
study
is
currently
being
conducted.
A
two­
generation
reproductive
toxicity
study
of
PFOS
showed
high
mortality
of
Fl
pups
at
doses
as
low
as
1.6
mg/
kg/
day.
The
results
of
the
APFO
study
will
be
important
to
determine
whether
a
similar
effect
is
observed.

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).
The
compound
(at
300
ppm)
has
also
been
reported
to
be
carcinogenic
toward
the
liver
and
pancreas
of
male
CD
rats.

The
mechanism(
s)
of
APFO
tumorigenesis
is
not
clearly
understood.
Available
data
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.
The
liver
carcinogenicity/
toxicity
of
APFO
appear
to
be
related
to
induction
of
peroxisome
proliferation
following
binding
to
the
peroxisome
proliferation
activation
receptor
a
(PPAR
a)
in
the
liver.
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
PPARo
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.

As
the
mechanisms
of
carcinogenic
action
of
APFO
have
not
been
fully
elucidated,
it
is
assumed
that
the
tumors
induced
in
rats
are
relevant
to
humans.
Review
of
available
mechanistic
data
of
other
drugs
and
chemicals
that
induce
LCT
in
animals
has
led
a
workshop
panel
to
conclude
that
all
but
two
modes
of
induction
of
the
luteinizing
hormone
(LH),
"dopamine
agonism"
and
"GnRH
agonism",
are
considered
to
be
relevant
to
humans,
and
that
the
possibility
of
induction
of
Leydig
cell
adenoma
in
humans
by
specific
agents
with
other
modes
of
action
cannot
be
ruled
out
despite
the
rarity
of
LCT
in
humans.
At
present,
there
is
no
evidence
that
the
induction
of
LCT
by
APFO
is
via
the
"dopamine
agonism"
or
"G&
H
agonism"
mode
of
action.
It
is
recognized
that
there
are
quantitative
differences
in
certain
biological
parameters
between
rats
and
humans.
However,
the
principal
cell
control
mechanisms
appear
similar,
and
the
difference
in
carcinogenic
response
is
probably
quantitative.
As
binding
to
the
PPARa
appears
to
be
the
critical
event
leading
to
hormonal
imbalance
and
APFO
tumorigenesis,
and
the
level
of
PPARcl
in
human
livers
is
lower
than
that
in
rodent
liver,
it
appears
that
humans
may
be
less
sensitive
than
rodents
in
the
development
of
LCT,
mammary
gland
tumors,
or
liver
neoplasms.
1.0
Chemical
Identity
Chemical
Name:
Perfluorooctanoic
Acid
Molecular
formula:
C8
H
F1.5
02
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
l]

Ammonium
Salt
Sodium
Salt
Potassium
Salt
Silver
Salt
(X
=
OM+;
M
=
NH4)
(X
=
OM+;
M
=
Na)
(X=
OM+;
M=
K)
(X
=
OM+;
M
=
Ag)
[3825­
26­
l]
[335­
95­
51
[2395­
00­
g]
[335­
93­
31
Acid
Fluoride
(X
=
F)
[335­
66­
O]

Methyl
Ester
Ethyl
Ester
(X
=
CH3)
[376­
27­
21
(X
=
CH2­
CH3)
[3
108­
24­
51
Synonyms
:
I­
Octanoic
acid,
2
,,,
233
>
4
>
4
>
5
>,
5
6
>
6
3
7
3
77
8
>
8,8­
pentadecafluoroPFOA
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
produced
by
the
telomerization
of
tetrafluoroethylene.
The
physical
chemical
properties
noted
below
are
for
the
free
acid,
unless
otherwise
stated.
The
data
for
the
free
acid,
pentadecafluorooctanoic
acid
[335­
67­
l],
is
the
most
complete.
The
reported
vapor
pressure
of
10
mm
Hg
appears
high,
but
is
consistent
with
other
perfluorinated
compounds
with
similar
boiling
points.
The
free
acid
is
expected
to
completely
dissociate
in
water.

Determination
of
the
vapor
pressure
of
APFO
is
problematic.
For
APFO,
the
recently
reported
vapor
pressure
of
<
1
x
IOE­
5
(3M
Environmental
Laboratory,
1993)
seems
too
low
for
a
material
that
sublimes
as
the
ammonium
salt.
This
study
measured
the
water
solubility
of
APFO
to
be
>
10%.
It
was
noted
in
an
earlier
study
that
concentrations
of
20
g/
L
"gelled"
(3M
6
Company,
1979).
The
partition
coefficient
was
reported
in
these
early
studies
of
5.
Another
calculated
value,
­0.9,
might
not
be
accurate
due
to
the
method
used
(Hansch
and
Leo
1979).
The
formation
of
an
emulsified
layer
between
the
octanol
and
water
surface
interface
would
make
determination
of
log
P
difficult.

The
available
physicochemical
properties
for
the
PFOA
free
acid
are:

MW:
414
(Beilstein,
1975)
MP:
45
­
50
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
AR­
226
473)
pH
(lg/
L):
2.6
(MSDS
Merck)

Due
to
the
surface­
active
properties
of
PFOA,
and
the
test
protocol
for
the
OECD
method,
PFOA
is
anticipated
to
form
multiple
layers
in
octanol/
water,
much
like
those
observed
for
PFOS.
Therefore,
an
n­
octanollwater
partition
coefficient
cannot
be
determined.
Water
solubility
has
been
reported
for
PFOA,
but
it
is
unclear
whether
these
values
are
for
a
microdispersion
of
micelles,
rather
than
true
solubility.
Several
reports
note
that
PFOA
salts
self­
associate
as
micelles
at
higher
concentrations.
(Simister,
1992;
Calfours,
1985;
Edwards,
1997).
In
aqueous
solutions,
micelles
partition
between
the
air
/
water
interface
on
the
surface.

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
(Beilstein
1975).
The
ammonium
salt
sublimes
at
130°
C
(USEPA
AR­
226
473).

The
physicochemical
properties
of
PFOA
and
its
derivatives
are
summarized
in
Table
1.

J3
Corrected
4/
l
Y2002
Table
1.
Reported
Physicochemical
Properties
2.0
Production
of
PFOA
and
its
Salts
PFOA
is
commercially
manufactured
by
two
major
alternative
processes:
1)
the
Simons
ElectroChemical
Fluorination
(ECF)
process
or
2)
the
telomerization
process.

In
the
ECF
process,
an
electric
current
is
passed
through
a
solution
of
anhydrous
hydrogen
fluoride
and
an
organic
feedstock
of
octanoic
acid
or
a
derivative.
The
ECF
process
replaces
the
carbon­
hydrogen
bonds
on
molecules
of
the
organic
feedstock
with
carbon­
fluorine
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)
perfluorooctanonyl
fluoride
(PFOF),
along
with
a
variable
mixture
of
byproducts
and
impurities.
The
output
of
the
ECF
process
is
not
a
pure
chemical,
but
instead
a
mixture
of
isomers
and
homologues
including
higher
and
lower
straight­
chain
homologues;
branched­
chain
perfluoroalkyl
fluorides
of
various
chain
lengths;
straight­
chain,
branched,
and
cyclic
perfluroalkanes
and
ethers;
and
other
byproducts
(3M
Company,
2000a).
After
disposal
or
recovery
of
some
of
the
byproducts
and
impurities,
the
acid
fluoride
is
base
hydrolyzed
in
batch
reactors
to
yield
PFOA.
The
PFOA
salts
are
synthesized
by
base
neutralization
of
the
acid
to
the
salt
in
a
separate
reactor
(3M
Company,
2000b).

In
the
telomerization
process,
tetrafluoroethylene
is
reacted
with
other
fluorine­
bearing
chemicals
to
yield
fluorinated
carboxylic
acids.
This
process
yields
pure
straight­
chain
acids
with
an
even
number
of
carbon
atoms.
Distillation
can
be
used
to
obtain
pure
components
(ECT,
1994).
Commercial
products
manufactured
through
the
telomerization
process
are
generally
mixtures
of
perfluorinated
compounds
with
even
carbon
numbers
(Renner,
2001).

3M
Company
is
the
largest
manufacturer
and
importer
of
PFOA
and
its
salts
in
the
United
States.
3M
has
characterized
its
manufacture
of
PFOA
and
its
ammonium
and
sodium
salts
in
1997
at
less
than
500,000
kg
per
year,
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
8
PFOA
is
consumed
in
the
manufacture
of
the
ammonium
or
sodium
salts.
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.

Industry
participants
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
(FMG,
2001).
USEPA
has
not
located
information
that
would
contradict
this
claim.
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.

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.

In
terms
of
on­
going
production,
3M
has
not
committed
publicly
to
a
complete
phase­
out
of
PFOA
and
PFOA­
related
chemicals
as
it
has
for
PFOS
and
PFOS­
related
chemicals.
However,
3M
has
indicated
that
it
is
phasing
out
certain
FLUORAD
Brand
specialty
materials
that
contain
PFOA
and
its
salts
such
as
FC­
26,
FC­
118
and
FC­
143,
FX­
1001
and
others
(3M
Company,
2OOOc).

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)
.
PCR
Inc.
(United
States)
.
Atofina
(France)
.
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
lnstitute
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
as
a
replacement
chemical
for
PFOS
in
other
applications
in
which
PFOA
is
not
currently
used.

10
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
"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
H202
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
Fe203
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
(Fe203)
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/
l­
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
Wmin.
(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
all~
yldimethylbenzylammonium
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
11
it
is
very
unlikely
that
these
fluorochemicals
were
removed
by
bubbling
air
through
water
due
to
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.

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.
I).
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
l/
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
HPLCYMSD
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
12
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
(HPLCYMS).

During
the
study,
samples
were
prepared
and
examined
at
six
different
pH
levels
from
1.5
to
11
.O
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
1
I)
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
To
determine
the
potential
for
bioaccumulation,
Fathead
minnows
were
exposed
to
25
mg/
l
APFO
for
13
days
(Howell
et
al.,
1995).
After
13
days
exposure,
the
fish
were
then
removed
from
APFO
contaminated
water
and
analyzed
for
depuration
over
15
days.
After
192
and
3
12
hours
exposure
to
APFO
contaminated
water,
the
average
concentration
of
APFO
in
fish
tissue
was
44.7
and
46.7
pg/
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
pg/
g
ww
and
by
96
hours
post­
exposure,
the
concentration
had
decreased
to
approximately
8
pg/
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.

Vraspir
(1979)
conducted
a
study
to
determine
if
bluegill
sunfish
bioaccumulate
fluorochemicals
from
the
3M
Decatur
plant.
Two
lots
of
30
Iish
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
CC/
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
13
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
"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.
Perfluorooctanoic
acid
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
Combustion
For
1997,3M
estimated
1950
pounds
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
Company,
2000a,
b).
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
Company,
2000a,
b).
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).

Canadian
research
has
stated
that
the
thermolysis
of
fluoropolymers,
e.
g.,
Teflon,
Kel­
F,
can
liberate
small
quantities
of
polycarboxylic
acids,
which
include
PFOA
(Ellis
et
al.,
2001).
This
information
was
insufficient
to
estimate
potential
yields.

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
1
4
1
4
following
APFO
water
releases
per
location:
Washington
Works
WV,
55,000
lbs;
Parlin
NJ,
300
lbs.;
Spruance
VA,
150
lbs.;
Chambers
Works
NJ,
9500
lbs.
(DuPont,
2000).

DuPont
measured
and
modeled
the
following
APFO
concentrations
at
its
sites:
Washington
Works
WV:
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
S/
l
about
50%
of
the
time
during
the
year,
and
likely
to
exceed
IO
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
0.3
13
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.

15
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
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.

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,
2001~).
PFOA
was
found
in
a
few
mink
livers
from
Massachusetts
at
a
concentration
range
of
~18
to
108
rig/
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
nglg,
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
five
downstream
sampling
locations,
the
two
closest
to
the
3M
facility
had
PFOA
surface
water
concentrations
significantly
greater
than
the
two
upstream
sites
(means
of
19OOug/
L
and
1024
ug/
L,
vs.
0.008
(est.)
and
0.028
ug/
L);
the
nearest
three
locations
had
sediment
concentrations
significantly
greater
than
the
upstream
sites
(wet
wt.
means
1855
ug/
kg,
892
ug/
kg,
238
ug/
kg
vs.
O.
O8(
est.)
and
O.
O9(
est.)).
Clam
and
fish
samples
were
collected
at
two
locations,
one
upstream
and
one
downstream
of
the
3M
facility.
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
average
PFOA
concentration
in
clams
at
the
upstream
location
was
4.38
ug/
kg;
that
for
the
downstream
location
was
8.42
ug/
kg.

2.4
Human
Biomonitoring
Table
1
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
16
levels
in
workers
exposed
to
PFOA
occupationally.
The
data
are
currently
limited
to
those
discussed
below.

3M
has
offered
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,
Minnesota;
Decatur,
Alabama;
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
I998
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,
and
the
mean
serum
PFOA
level
was
6.4
ppm
(range
=
0.1
­
8
1.3
ppm)
(Olsen
et
al.,
1998).
Only
74
employees
participated
in
the
1997
surveillance.
The
total
number
of
employees
working
at
the
plant
was
not
reported.

At
the
Decatur
plant,
263
of
500
employees
participated
in
2000
(Olsen
et
al.,
2001d).
The
mean
serum
PFOA
level
was
1.78
ppm.
This
was
slightly
higher
than
the
mean
in
1998
(1.54
ppm).
In
2000,
5
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.

Serum
PFOA
levels
at
the
Antwerp
plant
have
been
lower
than
at
Decatur
or
Cottage
Grove,
and
have
decreased
slightly
since
1995
(Olsen
et
al.,
2001e).
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,
and
the
highest
serum
level
reported
was
7.04
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.,
2OOlf).
The
mean
PFOA
serum
level
was
0.106
ppm
(range
0.008
­
0.668
Ppd.

Data
on
PFOA
levels
in
the
general
population
are
very
limited.
They
are
very
recent
and
are
only
available
on
small
cohorts.
The
mean
serum
PFOA
levels
are
much
lower
in
the
general
population
than
in
workers
exposed
to
PFOA.

1
7
1
7
Pooled
blood
samples
from
U.
S.
blood
banks
indicate
mean
PFOA
levels
of
3
to
17
ppb
(3M
Company,
1999a,
1999b).
The
highest
pooled
sample
reported
was
22
ppb.
Samples
were
collected
in
1998
and
1999.
These
data
provide
a
very
preliminary
view
of
the
PFOA
levels
that
may
be
present
in
the
U.
S.
general
population.
However,
it
cannot
be
assumed
that
these
levels
are
representative
of
the
U.
S.
population
for
several
reasons:
1)
blood
donors
are
not
necessarily
representative
of
the
U.
S.
population,
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.

Preliminary
data
on
individual
blood
samples
have
recently
been
reported
(Olsen
et
al.,
2001b,
2001~).
Blood
samples
from
652
U.
S.
adult
blood
donors,
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.
The
mean
serum
PFOA
level
was
5.6
ppb.
The
range
was
<lower
limit
of
quantitation
(LLOQ
=
1.92
or
2.11)
to
52.3
ppb.
Blood
samples
from
U.
S.
children
have
also
been
analyzed
for
serum
PFOA.
A
sample
of
599
children,
ages
2­
12
years
old,
participating
in
a
study
of
group
A
streptococcal
infections,
revealed
a
mean
PFOA
serum
level
of
5.6
ppb.
The
range
was
<LLOQ
to
56.1
ppb.
The
LLOQ
was
I
.92
or
2.88.
The
samples
were
collected
from
equal
numbers
of
male
and
female
children
residing
in
23
states.
The
samples
in
both
of
these
studies
were
analyzed
using
high­
pressure
liquid
chromatography/
electrospray
tandem
mass
spectrometry
(HPLCYESMSMS).
These
data
are
only
preliminary
and
have
not
completed
quality
assurance
procedures.

In
another
study,
the
PFOA
concentration
was
analyzed
in
human
sera
and
liver
samples
(Olsen
et
al.,
2001g).
Thirty­
one
donor
samples
were
obtained
from
16
males
and
15
females
over
an
18month
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
1IAM
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
rations
were
provided
because
more
than
90%
of
the
individual
liver
samples
were
<LOQ.

Serum
PFOA
levels
in
corporate
staff
and
managers
at
a
3M
plant
in
St.
Paul,
Minnesota,
where
occupational
exposure
to
PFOA
should
not
have
occurred,
were
reported
(3M
Company,
1999a).

18
Four
of
3
1
employees
had
serum
PFOA
levels
greater
than
the
detection
limit
of
10
ppb.
The
mean
for
these
employees
was
12.5
ppb.
Table
1.
SERUM
PFOA
LEVELS
IN
HUMAN
POPULATIONS
Occupational
Exposures
(nnm)

Plant
Arithmetic
Mean
Range
Geometric
95%
Mean
Confidence
Interval
Cottage
Grove
Plant
1997
(n
=
74)
6.4
0.1
­
81.3
*
*

1995
(n
=
80)
6.8
o.
o­
114.1
*
*

1993
(n=
111)
5.0
0.0
­
80.0
*
*

Decatur
Plant
2000
(n
=
263)
1.78
0.04
­
12.70
1.13
0.99
­
1.30
1998
(n
=
126)
1.54
0.02
­
6.76
0.90
0.72
­
1.12
1997
(n
=
84)
1.57
not
reported
*
*

1995
(n
=
90)
1.46
not
reported
*
*

Antwerp
Plant
2000
(n
=
2.58)
0.84
0.01
­
7.04
0.33
0.27
­
0.40
1995
(n
=
93)
1.13
O.
OO­
13.2
*
*

Building
236
2000
(n
=
45)
0.106
0.008
­
0.668
0.053
0.037
­
0.076
General
Population
Exposures
(ppb)
Source
1
Arithmetic
Mean
1
Range
Pooled
samples
Commercial
sources
of
blood,
1999
3
1
­
13
(n
=
35
lots)
Blood
Banks
(n
=
1
S),
1998
­340­
680
donors
17**
12
­22
Individual
samples
American
Red
Cross
blood
banks,
2000
5.6
4.27
­
52.3
(n
=
652)
Children,
1995
5.6
4.27
­
56.1
(n=
599)
3M
Corporate
managers/
staff
St.
Paul,
MN,
1998
12.5"""
not
reported
(n=
31)
*Geometric
mean
and
95%
confidence
intervals
were
not
included
in
the
reports.
*
*PFOA
detected
in
about
l/
3
of
the
pooled
samples
but
quantifiable
in
only
2
***
only
4
employees
were
above
the
detection
limit
of
10
ppb
19
3.0
Human
Health
Hazards
3.1.
Metabolism
and
Pharmacokinetics
3.1.1
Half­
life
in
Humans
In
order
to
determine
the
half­
life
of
PFOA,
a
group
of
retirees
(n
=
20)
volunteered
to
participate
in
a
5­
year
half­
life
study
in
which
serum
samples
will
be
drawn
every
6
months
(Burris
et
al.,
2000).
The
only
other
data
available
on
the
half­
life
of
PFOA
is
from
a
1980
study
in
which
it
was
estimated
to
be
approximately
1
year;
however,
this
analysis
was
based
on
total
organic
fluorine
in
blood
serum.

Twenty­
seven
retirees,
age
55
to
74
years,
volunteered
to
participate
in
this
half­
life
study.
PFOA
levels
in
this
group
ranged
from
0.1
to
3.1
ppm.
Most
of
the
retirees
were
employed
at
the
Decatur,
Alabama
plant
for
an
average
of
28
years.
The
number
of
years
since
retirement
varied
greatly
among
the
participants.
The
average
length
of
time
between
retirement
and
the
start
of
the
study
was
30
months
(2.5
years)
but
ranged
from
5
to
130
months
(­
.5
to
10
years).
There
were
3
collection
periods
during
which
serum
PFOA
samples
were
collected
and
analyzed:
November
1998,
June
1999,
and
November
1999.
Cottage
Grove
employees,
where
PFOA
exposure
was
much
higher,
have
only
participated
in
2
sampling
periods;
therefore,
they
were
not
included
in
this
analysis.

Half­
lives
were
calculated
using
a
one­
compartment
model.
A
log­
linear
relationship
(slope
=
&(
2.303))
was
used
to
estimate
the
half­
life.
The
half­
life
was
calculated
after
the
elimination
constant
was
determined,
using
the
relationship:
ti/
2
=
0.693&
Only
those
retirees
who
fit
the
linear
one
compartmental
model
(r2
$
0.6)
for
PFOA
were
included
in
the
analyses.
If
3
data
points
were
not
available
for
any
of
the
subjects
and
if
there
was
a
lack
of
fit
to
the
model,
that
retiree
was
not
included
in
the
analysis.
Twenty
participants
met
these
requirements.
The
median
serum
half­
life
of
PFOA
was
344
days,
with
a
range
of
109
to
1308
days.
The
two
highest
half­
life
calculations
were
for
the
2
female
retirees
who
participated
in
this
study
(654
and
1308
days).
It
should
be
noted
that
the
difference
in
PFOA
serum
levels
between
retirees
was
quite
large
(0.1
­
3.1
ppm).
It
was
not
specifically
stated
in
the
report;
however,
based
on
a
statement
in
the
report,
it
is
assumed
that
the
2
female
retirees
did
not
have
the
highest
PFOA
serum
levels.

For
most
of
the
participants
not
included
in
the
analysis,
the
second
measurement
was
higher
than
the
first.
Therefore,
the
data
did
not
fit
the
model
and
they
were
excluded.
Although
this
may
justify
not
including
those
participants
in
the
analysis,
it
is
an
indication
of
the
many
limitations
of
the
data.
It
is
stated
in
the
report
that
neither
age
nor
number
of
months
retired
was
associated
with
the
serum
PFOA
half­
life
calculations;
however,
this
statement
is
not
supported
with
any
data
in
the
report.
In
addition,
no
individual
data
were
provided
in
the
report
and
the
relationship
between
number
of
years
exposed
in
the
workplace
and
PFOA
levels
and
half­
life
were
excluded.
Also,
elimination
of
PFOA
occurs
via
urine
and
feces;
however,
these
20
measurements
were
not
collected.
Therefore,
it
cannot
be
determined
whether
the
half­
life
suggested
by
the
preliminary
results
reported
here
represents
a
true
elimination
half­
life
from
the
body.
Finally,
the
effect
of
continued
non­
occupational,
low­
level
exposure
on
the
half­
life
is
unknown.

The
data
presented
above
provide
a
very
rough
estimate
of
the
plasma
half­
life
of
PFOA.
It
does
not
provide
an
elimination
rate.
In
addition,
these
data
do
not
provide
any
information
about
the
distribution
of
PFOA
in
the
body.
Without
the
individual
data
or
supporting
information,
the
statement
that
time
between
retirement
and
entry
into
the
study
does
not
affect
the
half­
life
calculation
is
highly
suspect.
One
would
expect
age,
length
of
exposure,
and
time
elapsed
since
occupational
exposure
to
affect
PFOA
serum
levels.
Since
these
data
were
not
provided
in
the
report
and
since
only
3
data
points
have
been
calculated
to
date,
one
can
only
estimate
that
the
half­
life
of
PFOA
is
between
1
and
3.5
years.
These
data
provide
evidence
of
the
potential
to
bioaccumulate
PFOA
in
humans.
In
addition,
these
preliminary
data
suggest
that
gender
plays
a
role
in
the
half­
life.

3.1.2
Absorption
Studies
in
Animals
APFO
is
well
absorbed
following
oral
and
inhalation
exposure,
and
to
a
lesser
extent
following
dermal
exposure.
In
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
inmbation
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
.O
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
hi­/
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
mgikg
resulted
in
elevated
blood
organofluorine
levels
which
increased
in
a
dose­
related
manner
(Kennedy,
1985).

O'Malley
and
Ebbins
(198
1)
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
21
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%
(O/
4)
in
the
100
mg/
kg
group.

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
be
absorbed
dermally,
work
practices
were
changed
and
engineering
controls
were
adopted
that
reduced
dermal
exposures
(Gilliland,
1992).

A
t­
butyl
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
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.
Major
sex­
related
differences
in
the
disposition
of
PFOA
have
been
observed.

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
63
ppm
and
females
75
ppm.
Liver
levels
were
9
and
10
ppm
for
males
and
females,
respectively.

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.

22
Twenty­
four
hours
after
oral
administration
of
APFO
(2
mg
APFO
in
2
ml
aqueous
solution
by
stomach
intnbation),
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.

Vanden
Heuvel
et
al.
(199
1)
investigated
the
di,
sFosition
of
PFOA
in
perfused
male
rat
liver.
Approximately
11%
of
the
cumulative
dose
of
C­
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/
lOOg)
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/
lOOg)
to
newly
weaned
rats
(18
male,
18
female).
After
the
single
dose,
samples
were
collected
for
PFOA
determination
12,
24­
168
(at
24
h
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
h
was
estimated
in
the
liver
during
the
first
week.
In
the
males,
the
half­
life
in
liver
was
2
10
h.
Although
PFOA
was
retained
by
the
liver,
it
was
not
found
in
the
lipid
fraction.
In
the
kidney,
the
half­
life
was
145
h
and
130
h
in
females
and
males,
respectively.
In
the
spleen,
the
half­
life
was
73
h
and
170
h
in
23
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
5
1.65
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
23.7
1
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.

Vanden
Heuvel
et
al.
(1992)
demonstrated
that
PFOA
covalently
binds
to
proteins
in
the
liver,
plasma,
and
testes
of
rats
in
vivo.
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
in
vitro
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
in
vitro
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/
lOOg)
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+­
l
1.2
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.

24
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
in
vivo.
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,
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+­
l
.O).
Excretion
in
urine
was
4
1%
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+­
l
.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
IO­
fold
suggests
that
there
is
enterohepatic
circulation
of
APFO.

The
disposition
of
PFOA
(tetrabutyl
ammonium
salt
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,
nondetectable
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.
(
199
1)
investigated
the
in
vivo
metabolism
of
PFOA
in
rats
administered
14C­
PFOA
(9.4
umol/
kg,
i.
p.).
Pooled
daily
urine
samples
(O­
4
days
post­
treatment)
and
bile
25
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
in
vivo.

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
intraperotoneal
PFOA
dose
(50
mg/
kg)
in
male
and
female
rats.

3.1.5
Elimination
Studies
in
Animals
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
Heuvel
et
al.,
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
9
1%
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
renalligated
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
(CO.
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
26
32
recovered
in
the
urine.
The
quantity
of
nonionic
fluorine
recovered
in
the
urine
increased
to
6
1%
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
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
[
I­
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
0.2
1
ml/
mm.
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
lo20
and
unlabeled
APFO
(80­
90%).
Five
percent
mannitol
in
isotonic
saline
was
infused
at
a
rate
of
0.08
1
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+­
O.
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+­
l
.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/
lOOg.
Net
APFO
excretion
was
reduced
from
4.6
to
0.13
ug/
min/
lOOg.
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/
lOOg,
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
PFO
elimination
to
11.8%.
It
is
concluded
that
the
female
rat
possesses
an
active
secretory
mechanism
which
rapidly
eliminates
APFO
from
the
body.
This
27
33
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.
PFOA
was
administered
by
gavage
to
48
newlyweaned
animals
at
0,
3,
10,
and
30
mg/
kg
(in
0.9%
NaCl,
0.5ml/
lOOg)
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
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
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.
From
these
dogs,
plasma
PFOA
half­
lives
were
determined.
There
was
no
difference
between
the
renal
clearances
of
the
male
and
female
dogs
either
before
or
after
probenecid.
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
half­
lives
of
PFOA
were
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
intraperotoneal
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+­
l
9%
and
68+­
14%,
respectively).
The
castrated
males
without
28
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
(>
l­
fold)
the
elimination
of
14'
­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
excretion
of
PFOA.
In
female
rats,
neither
ovariectomy
nor
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
carbon­
14
after
oral
administration
of
a
single
dose
of
ammonium
`4C­
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
carbon­
14
content.
Essentially
all
of
the
carbon­
14
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
ammonium
[
14lperfluorooctanoate
(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
9.8fold
the
mean
percentage
of
dose
eliminated
in
feces
by
untreated
rats
(4.4­
t­
1
.O).
Excretion
in
urine
was
41%
for
treated
rats
and
67%
for
untreated
rats.

3.2
Epidemiology
Studies
3.2.1
Mortality
Study
A
retrospective
cohort
mortality
study
was
performed
on
employees
at
the
Cottage
Grove,
Minnesota
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
29
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.

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
employmen
in
the
Chemical
Division
and
prostate
cancer
mortality.
Based
on
the
results
of
proportional
hazard
models,
the
relative
risk
for
a
l­
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).

30
36
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
=
1.3
1,
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),
1
from
pancreatic
cancer
(SMR
=
1.34),
and
1
from
prostate
cancer
(SMR
=
1.30).
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%
CT
=
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
(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
<
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
3
1
3
1
37
disease
than
the
general
population.
A
dose­
response
relationship
was
not
observed
with
years
of
exposure.

The
slight
excess
in
bladder
cancer
in
the
cohort
as
a
whole
should
be
noted,
especially
given
the
bladder
cancer
mortality
experience
at
3M's
Decatur
plant,
which
produces
mostly
PFOS.
Bladder
cancer
mortality
was
4
times
higher
in
workers
with
high
PFOS
exposure
jobs
at
3M's
Decatur,
Alabama
plant
than
the
general
population
(SMR
=
4.81,
95%
CI
=
0.99
­
14.06)
(Alexander,
2OOlb).
Three
deaths
were
reported,
and
all
of
them
occurred
in
the
high
exposure
group.
Serum
PFOA
levels
in
workers
are
lower
at
the
Decatur
plant,
where
PFOA
is
used
as
an
elastomer
in
fluoropolymer
production
or
is
produced
as
a
by­
product,
than
at
Cottage
Grove;
however,
the
manufacture
of
PFOA
began
at
Decatur
in
1999.
Therefore,
PFOA
exposures
will
likely
increase
at
this
plant.
It
is
not
clear
whether
PFOA,
PFOS
or
some
other
chemical
may
be
responsible
for
the
bladder
cancer
deaths
observed
at
these
plants;
however,
follow
up
should
continue
in
an
effort
to
shed
some
light
on
this
observation.

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
authors
claim
that
the
newer
exposure
categories
are
more
accurate,
it
is
still
likely
that
exposure
misclassification
occurred.
Without
measured
exposures
(serum
PFOA
levels),
it
is
difficult
to
judge
the
reliability
of
the
exposure
categories
that
were
defined.
In
the
second
study,
the
chemical
plant
employees
were
sub­
divided
into
PFOA­
exposed
groups,
and
the
film
plant
employees
essentially
remained
in
the
"non­
exposed"
group.
This
was
an
effort
to
more
accurately
classify
exposures;
however,
these
new
categories
do
not
take
into
account
duration
of
exposure
or
length
of
employment.
Another
limitation
to
this
study
is
that
17
death
certificates
were
not
located
for
deceased
employees
and
therefore
were
not
included
in
the
study.
The
inclusion
or
exclusion
of
these
deaths
could
greatly
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.

3.2.2
Hormone
Study
Endocrine
effects
have
been
associated
with
PFOA
exposure
in
animals;
therefore,
2
crosssectional
studies
were
conducted
on
employees
of
a
plant
producing
PFOA
(Olsen,
et
al.,
1998a).
Medical
surveillance,
hormone
testing
and
PFOA
serum
levels
were
obtained
for
volunteer
workers
in
1993
(n
=
111)
and
1995
(n
=
SO).
Sixty­
eight
employees
were
common
to
both
sampling
periods.
In
1993,
the
range
of
PFOA
was
O­
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).

32
Employees
were
placed
into
4
exposure
categories
based
on
their
serum
PFOA
levels:
O­
l
ppm,
I­
<
10
ppm,
lo­
<
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
<
.OOl
in
1993,
and
r
=
.30,
p
<
.Ol
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
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
3
1
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
33
34
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.

In
1997,
voluntary
medical
surveillance
was
again
offered
to
employees
(Olsen,
et
al.,
1998b).
In
this
sampling
period,
cholecystokinin
(CCK)
levels
were
analyzed
in
74
employees
to
determine
if
they
were
positively
associated
with
serum
PFOA
levels.
CCK
levels
were
observed
because
research
has
suggested
that
pancreas
acinar
cell
adenomas
seen
in
rats
exposed
to
PFOA
may
be
the
result
of
increased
CCK
levels.
Seventeen
of
the
subjects
were
common
to
all
three
sampling
periods
(1993,
1995,
and
1997).
The
same
statistical
methods
were
used
in
this
study
period
as
used
in
1993
and
1995,
and
the
four
PFOA
exposure
categories
were
also
the
same.

The
mean
PFOA
serum
level
in
employees
participating
in
the
1997
study
period
was
6.4
ppm
(range
0.1
­
81.3
ppm).
The
mean
CCK
value
was
28.5
pg/
ml
(range
8.8
­
86.7
pg/
ml).
The
highest
CCK
values
were
reported
in
the
2
exposure
categories
less
than
10
ppm.
The
means
were
50%
higher
in
these
2
categories
than
in
the
categories
greater
than
10
ppm
(p
=
.06).
When
adjusted
for
potential
confounders,
multivariable
regression
models
indicated
a
weak
negative
association
between
CCK
and
PFOA;
however,
the
data
were
not
included
in
the
report.

The
following
explanations
may
indicate
why
this
study
failed
to
find
a
positive
association
between
PFOA
and
CCK
values:

.
It
is
possible
that
the
hepatocarcinogenic
effects
of
peroxisome
proliferators
in
rodents
do
not
act
the
same
biochemically
in
humans.
.
The
serum
PFOA
levels
observed
in
workers
may
have
been
too
low
to
detect
an
effect.
Effects
in
animals
were
observed
at
higher
doses
than
most
of
the
serum
levels
found
in
workers.
.
CCK
receptors
may
be
different
between
rats
and
humans.
Therefore,
the
monkey
may
be
a
more
appropriate
animal
model
to
study
the
pancreatic
effects
of
PFOA
in
humans.
.
The
involvement
of
CCK
in
the
initiation
or
promotion
of
pancreatic
cancer
is
controversial.
.
The
rat
may
not
be
an
appropriate
model
in
the
study
of
pancreatic
cancer
in
humans,
since
acinar
cell
malignancies,
induced
by
carcinogens
in
rats,
are
rare
in
humans.

The
same
methodological
issues
that
applied
to
the
study
in
1993
and
1995
apply
to
this
portion
of
the
study
as
well.

34
w
3.2.3
Cholesterol
Study
Based
on
animal
testing
which
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
(Gilliland
and
Mandel,
1996).
Forty­
eight
workers
who
were
exposed
to
PFOA
from
1985­
1989
were
included
in
the
study
(96%
participation
rate).
Sixtyfive
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),
serum
glutamyl
pyruvic
transaminase
(SGPT),
gamma
glutamyl
transferase
(GGT),
cholesterol,
low­
density
lipoproteins
(LDL),
and
high­
density
lipoproteins
(HDL).
All
of
the
participants
were
placed
into
five
categories
of
total
serum
fluorine
levels:
<I
ppm,
l­
3
ppm,
>3
­
10
ppm,
>I
0
­
15
ppm,
and
>
1.5
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
SGOT,
SGPT,
and
GGT.
However,
increases
in
SGOT
and
SGPT
occurred
with
increasing
total
serum
fluorine
levels
in
obese
workers
(BMI
=
35
kg/
m*).

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
35
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).

3.2.4
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.
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
3
1,
1998.
Based
on
work
history
records,
employees
were
placed
into
different
comparison
groups:
Group
A
consisted
of
all
film
and
chemical
plant
workers;
Group
B
had
employees
who
only
worked
in
either
the
film
or
chemical
plant;
Group
C
consisted
of
employees
who
worked
in
jobs
with
high
POSF
exposures;
and
Group
D
had
employees
who
worked
in
high
exposures
in
the
chemical
plant
for
10
years
or
more
prior
to
the
onset
of
the
study.
Film
plant
employees
were
considered
to
have
little
or
no
fluorochemical
exposure,
while
chemical
plant
employees
were
assumed
to
have
the
highest
exposures.

Ratios
of
observed
to
expected
episodes
of
care
were
calculated
for
each
plant.
Expected
numbers
were
based
on
3M's
employee
population
experience
using
indirect
standardization
techniques.
A
ratio
of
the
chemical
plant's
observed
to
expected
experience
divided
by
the
film
plant's
observed
to
expected
experience
was
calculated
to
provide
a
relative
risk
ratio
for
each
episode
of
care
(RREpC).
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
36
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
plant
employees
and
film
plant
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,
9.5%
CI
=
1.7
­
5.2).
Most
of
the
episodes
were
attributable
to
benign
colonic
polyps.
Similar
numbers
of
episodes
were
reported
in
film
and
chemical
plant
employees.

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

Chemical
plant
employees
in
the
high
exposure,
long­
term
employment
group
were
2
%
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
37
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
2464
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,
cornea1
opacity,
and
hypothermic
to
touch.

In
one
study
(Internat'l
Res
and
Dev
Corp.,
1978),
the
oral
LD50
values
for
Charles
River
CD
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)
<lo00
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).

38
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.

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,
l­
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
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
Sulmonella
­E.
colilmammalian­
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.
39
APFO
did
not
induce
chromosomal
aberrations
in
vitro
in
human
lymphocytes
when
tested
with
and
without
metabolic
activation
up
to
cytotoxic
concentrations
(Murli,
1996~;
NOTOX,
2000).

APFO
was
tested
twice
for
its
ability
to
induce
chromosomal
aberrations
in
CHO
cells
in
vitro.
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
1
OTI/,
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
in
vivo
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
Two
unpublished
2%
day
feeding
studies
were
performed
at
Industrial
Bio­
Test
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.

In
a
28day
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.
At
the
beginning
of
the
study
the
animals
averaged
88
grams
for
males
and
76
grams
for
females.
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.

40
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.
At
the
beginning
of
the
study
the
animals
averaged
88
grams
for
males
and
76
grams
for
females.
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
lOOO­
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
lOO­
ppm
groups
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(
l978b)
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
lOO­
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
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
lo­
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).

42
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/
lOO
ml
in
the
control).
The
authors
of
the
report
attribute
this
to
a
single
high
value
for
male
#7366
who
had
a
value
of
13
1.
The
other
three
monkeys
in
the
3­
mg/
kg/
day
groups
had
levels
of
112,
105,
and
120­
mg/
lOO
ml.
Glucose
levels
in
the
10
and
30
mg/
kg/
day
groups
were
104
and
122­
mg/
lOO
ml,
respectively,
after
one
month
of
treatment.
At
three
months
of
treatment,
glucose
levels
were
81,
96,
88,
and
66­
mg/
lOO
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
W/
l
in
the
control)
at
one
month,
which
persisted
in
the
one
surviving
male
(360
vs
85
1
W/
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
lo­
mg/
kg/
day
group
were
comparable
to
the
controls
at
3
months
(34
vs
31
II­
J/
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/
lOOml)
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
mgikglday
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
lOO­
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
mglkglday
groups.

The
levels
of
PFOA
in
the
serum
and
liver
are
presented
below.
Dose
Serum
(ppm)
Liver
(ppm)
Liver
total
(ug)

0
0
3
3
10
10
30
30
100
F
e
m
a
l
e
s
Males
Males
Females
Males
Females
ND
1
0.05
0.07
3
5
53
65
3
7
250
350
48
50
ND
ND
ND
ND
45
79
9
ND
600
ND
71
71
ND
10
ND
750
ND
ND
125
80
8000
7500
145
ND
60
125
4000
9000
ND
ND
100
325
6000
20000
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
0.56,
1.72,
5.64,
17.9,
and
63.5
mgikglday
in
males,
and
0.74,
2.3,
7.7,
22.36
and
76.47
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,
sex
and
group,
frozen
and
shipped
to
the
sponsor
for
analysis,
pooled
serum
samples.
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.

44
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
IOOO­
ppm
group
(362
g
vs
466
g
in
the
control
group).
Food
consumption
was
reduced
in
males
in
the
100,300
and
lOOO­
ppm
groups,
but
the
differences
were
not
statistically
significant.

Males
in
the
30,
100,300
and
lOOO­
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
lOOO­
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
IO­
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
lOOO­
ppm
dose
levels
had
increased
glucose
levels
(mg/
lOO
ml),
which
were
statistically
significant
at
all
but
the
lOO­
ppm
dose
level.
Reported
glucose
levels
were
121,
120,
136,
134,
143
and
135
mg/
lOO
ml
for
the
0,
lo,
30
100,300
and
1000
ppm
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/
lOO
ml
for
the
three
dose
groups,
respectively,
compared
to
16.2
mg/
lOO
ml
for
the
controls.
Alkaline
phosphatase
was
elevated
in
males
in
the
100,
300,
and
lOOO­
ppm
groups;
the
levels
were
147,204
and
2
12
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
1
OOO­
ppm
dose
level.
These
animals
had
enlarged
livers
that
showed
varying
degrees
of
surface
discoloration.
Neither
females
from
the
lOOO­
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
1
OOO­
ppm
groups
and
in
one
female
in
the
1
OOO­
ppm
group.
Compound­
related
liver
lesions
occurred
in
all
male
rats
in
the
100,
300
and
lOOO­
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
45
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
1
OOO­
ppm
dietary
level.

A
comparison
of
the
serum
levels
of
PFOA
is
shown
below.
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.

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
0.05,
0.47,
1.44,
and
4.97
mg/
kg/
day)
APFO
ad
Zibitum
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
dosedgroups
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
dosed­
group
were
sacrificed
following
4,
7,
or
13
weeks
of
treatment;
10
animals
per
dosedgroup
were
sacrificed
after
13
weeks
of
treatment
and
following
8
weeks
of
non­
treatment.
Serum
samples
collected
from
10
animals
per
dosed­
group
at
each
scheduled
sacrifice
during
treatment
and
from
5
animals
per
dosed­
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
dosed­
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
46
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
lOO­
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
l­
1
3
(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
l­
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
analysis
of
test
material
residue
showed
no
increase
in
serum
APFO
levels
over
the
course
of
treatment.
Statistically
significant
higher
hepatic
palmistry
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.
All
dosed
groups
exhibited
significant
increases
in
absolute
and
relative
liver
weights
and
hepatocellular
hypertrophy
were
observed
at
weeks
4,
7,
and
13,
compared
to
the
pair­
fed
control
group.
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
.O
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.

47
53
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
CD
3,6,
9,
12,
15,
and
20.
Animals
were
sacrificed
on
GD
20
by
cervical
dislocation
and
the
ovaries,
uteri,
and
contents
were
examined
for
the
number
of
corpora
lutea,
number
of
viable
and
non­
viable
fetuses,
number
of
resorption
sites,
and
number
of
implantation
sites.
Fetuses
were
weighed
and
sexed
and
subjected
to
external
gross
necropsy.
Approximately
one­
third
of
the
fetuses
were
fixed
in
Bouin's
solution
and
examined
for
visceral
abnormalities
by
free­
hand
sectioning.
The
remaining
fetuses
were
subjected
to
skeletal
examination
using
alizarin
red.

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

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

A
second
oral
prenatal
developmental
toxi,
city
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
48
(GD)
6­
l
8.
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
37'
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
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
13'"
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
13"
'
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
APFO
by
whole­
body
vapor
inhalation
to
0,
O.
1,
1,
10,
and
25
mg/
m3
6
hours/
day,
on
GD
6­
l
5.
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
2
1
and
observed
daily
for
abnormal
clinical
signs.
On
GD
2
1,
the
dams
were
sacrificed
by
cervical
dislocation
and
examined
for
any
gross
abnormalities,
liver
weights
were
recorded
and
the
reproductive
status
of
each
animal
was
evaluated.
The
ovaries,
uterus
and
contents
were
examined
for
the
number
of
corpora
lutea,
live
and
dead
fetuses,
resorptions
and
implantation
sites.
Pups
(live
and
dead)
were
counted,
weighed
and
sexed
and
examined
for
external,
visceral,
and
skeletal
alterations.
The
heads
of
all
control
and
high­
dosed
group
fetuses
were
examined
for
visceral
alterations
as
well
as
macro­
and
microscopic
evaluation
of
the
eyes.

For
trial
two,
in
which
the
dams
were
allowed
to
litter,
the
procedure
was
the
same
as
that
for
trial
one
up
to
GD
21.
Two
days
before
the
expected
day
of
parturition,
each
dam
was
housed
in
an
individual
cage.
The
date
of
parturition
was
noted
and
designated
Day
1
PP.
Dams
were
weighed
and
examined
for
clinical
signs
on
Days
1,7,
14,
and
22
PP.
On
Day
23
PP
all
dams
were
sacrificed.
Pups
were
counted,
weighed,
and
examined
for
external
alterations.
Each
pup
was
subsequently
weighed
and
inspected
for
adverse
clinical
signs
on
Days
4,
7,
14,
and
22
PP.
The
eyes
of
the
pups
were
also
examined
on
Days
15
and
17
PP
for
the
inhalation
portion
and
on
Days
27
and
3
1
PP
for
the
gavage
portion
of
the
study.
Pups
were
sacrificed
on
Day
35
PP
and
examined
for
visceral
and
skeletal
alterations.

Inhalation
Exuosure
Trial
One:

Treatment­
related
clinical
signs
of
maternal
toxicity
for
trial
one
(teratology)
occurred
at
10
and
25
mg/
m"
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,
was
indicated.

50
No
effects
were
observed
on
the
maintenance
of
pregnancy
or
the
incidence
of
resorptions.
Mean
fetal
body
weights
were
significantly
decreased
in
the
25mg/
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/
m",
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,
was
indicated.

Trial
Two:

Clinical
signs
of
maternal
toxicity
seen
at
10
and
25
mg/
m3
were
similar
in
type
and
incidence
as
those
described
for
trial
one.
Maternal
body
weight
gain
during
treatment
at
25
mg/
m3
was
less
than
controls,
although
the
difference
was
not
statistically
significant.
In
addition,
2
out
of
12
dams
died
during
treatment
at
25
mg/
m3.
No
other
treatment­
related
effects
were
reported,
nor
were
any
adverse
effects
noted
for
any
of
the
measurements
of
reproductive
performance.
Under
the
conditions
of
the
study,
a
NOAEL
and
LOAEL
for
maternal
toxicity
of
1
and
10
mg/
m3,
respectively,
were
indicated.

Signs
of
developmental
toxicity
in
this
group
consisted
of
statistically
significant
reductions
in
pup
body
weight
on
Day
1
PP
(6.1
g
at
25
mg/
m3
vs.
6.8
g
in
controls).
On
Days
4
and
22
PP,
pup
body
weights
continued
to
remain
lower
than
controls,
although
the
difference
was
not
statistically
significant
(Day
4
PP:
9.7
g
at
25
mg/
m3
vs.
10.3
in
controls;
Day
22
PP:
49.0
g
at
25
mg/
m3
vs.
50.1
in
controls).
No
significant
effects
were
reported
following
external
examination
of
the
pups
or
with
ophthalmoscopic
examination
of
the
eyes.
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
Carcinogenicity
Studies
in
Animals
3.7.1
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
FC­
143
for
two
years.
Groups
of
15
additional
rats
per
sex
were
fed
0,
or
300
ppm
FC­
143
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
l
8
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
O%,
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
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
(a
clinical
sign)
and
ovarian
tubular
hyperplasia
(which
is
reversible),
the
LOAEL
for
female
rats
is
30
ppm.]

52
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
(PC
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
O%,
4%
and
14%,
respectively;
the
respective
incidences
of
mammary
fibroadenoma
in
the
female
groups
were
22%,
42%
and
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.

In
a
follow­
up
2­
year
dietary
study
(300
ppm)
in
male
Sprague­
Dawley
(CD)
rats,
APFO
was
found
to
induce
liver
tumors
and
pancreatic
acinar
cell
tumors
in
addition
to
Leydig
cell
tumors;
however,
details
on
the
study
design
and
tumor
incidence
were
not
reported
(Cook
et
al.,
1994).
APFO
has
also
been
shown
to
promote
liver
carcinogenesis
in
rodents
(Abdellatif
et
al.,
199
1;
Nilsson
et
al.,
1991).

3.7.2
Mode
of
Action
Studies
The
mechanisms
of
toxicological/
carcinogenic
action
of
APFO
are
not
clearly
understood.
Short­
term
genotoxicity
assays
suggest
that
APFO
is
not
a
DNA­
reactive
compound;
it
is
nonmutagenic
in
the
Ames
test
using
five
strains
of
Salmonella
typhimurium,
or
in
an
assay
with
Saccharomyces
cerevisiae
(Griffith
and
Long,
1980).
Available
data
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.

3.7.2.1
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.,
Ileda
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
53
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.

3.7.2.2
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
l3,
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
(r.
(TGF
a),
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
P­
oxidation
and
cytochrome
P­
450
monooxygenases
by
binding
to
the
peroxisome
proliferation
activation
receptor
c1
(PPAR
a;
a
subfamily
of
steroid
hormone
receptors).
It
is
believed
that
APFO
induces
cytochrome
P450
XIX
(aromatase)
by
binding
to
and
activating
the
PPARa.

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,
which
is
synthesized
and
secreted
by
the
Leydig
cells,
is
regulated
by
LH;
testosterone
and
LH
form
a
closed­
loop
feedback
system
in
the
HPT
axis.
In
order
to
maintain
adequate
testosterone
plasma
levels,
reduced
testosterone
levels
(caused
by
increased
aromatase
activity)
are
expected
to
lead
to
increased
LH
levels
through
the
negative
54
60
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).

3.7.2.3
Mammary
Gland
Tumors
Estradiol
has
also
been
shown
to
stimulate
the
secretion
of
TGF
c1
by
mammary
epithelial
cells
and
the
overexpression
of
TGF
c1
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
tibroadenomas
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).

3.7.2.4
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
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.

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.

5.5
61
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
l
996.
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
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
>lOO
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
2
(organized
by
test
substance)
and
3
(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
LC5Os
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
>lOOO
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
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
(.
O.
lslO
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.

57
63
Table
2
Summary
of
Acute
Ecological
Toxicity
Data
for
APFO
(grouped
by
test
substance)
Test
Organism
Duration
Value
Reference
(w/
U*
Test
Sample:
APFO
ammonium
salt
Fathead
minnow
(Pimephales
promelas)
9
6
­h
LC50
7
0
3M
Company,
1974a
96­
h
LC50
766
3M
Company,
1980a
Bluegill
sunfish
(Lepomis
machrochirus)

Water
flea
(Daphnia
magna)
96­
h
LC50
96­
h
LC50
96­
h
LC50
96­
h
LC50
48­
h
EC50
48­
h
EC50
301
3M
Company,
1987~
740
Ward
et
al.,,
1995
>
420
3M
Company,
1978
569
3M
Company,
1978
126
3M
Environmental
Laboratory,
1982
>
1000
3M
Environmental
Laboratory,
1982
Bacteria
(Photobacterium
phosphoreum)

Fathead
minnow
(Pimephales
promelas)

58
Water
flea
(Daphnia
magna)
48­
h
EC50
T.
R.
Wilbury
Laboratories,
Inc.,
1996b
Green
algae
(Selenastrum
capricornutum)
96­
h
EC50
90
T.
R.
Wilbury
Laboratories,
Inc.,
1995
I
Test
Sample:
APFO
(44%)
in
27.9%
water
and
27.2%
isopropanol
Fathead
minnow
(Pimephales
promelas)
196­
h
EC50
1391
T.
R.
Wilbury
Laboratories,
Inc.,

Fathead
minnow
(Pimephales
promelas)
96­
h
EC50
422
I
I
Test
Sample:
APFO
(44%)
in
27.9%
water
and
27.2%
Water
flea
(Duphnia
magna)
48­
h
EC50
41
Water
flea
(Duphnia
magna)
48­
h
EC50
39
Green
algae
(Seienastrum
capricornutum)
96­
h
EC50
2.1
Green
algae
(Selenastrum
capricornutum)
96­
h
EC50
3.6
Green
algae
(Selenastrum
capricornutum)
96­
h
EC50
1.2
*Values
were
adjusted
to
represent
100%
active
ingredient.
*These
values
may
be
inconsistent
due
to
different
diets
tested.
T.
R.
Wilbury
Laboratories,
1
9
9
5
isopropanol
IWard
et
al..
1995
IWard
et
al..
1995
IWard
et
al..
1995
IWard
et
al.,
1995
/Ward
et
al.,
1995
Inc.,

­I
5
9
Table
3
Summary
of
Ecological
Toxicity
Data
for
APFO
(grouped
by
species)
Test
Organism
Value
Reference
Duration
OWL)

Fathead
minnow
(Pimephalespromelas)
96­
h
LC50
70'
96­
h
LC50
766"
96­
h
LC.
50
301B
3M
Company,
1974a
3M
Company,
1980a
3M
Company,
1987~

196­
h
LC50
1>
500D
IEnviroSvstems,
Inc.,
1990a
I
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
96­
h
EC50
>
100B
EG&
G
Bionomics
Aquatic
Toxicology
Laboratory,
1978
391o
T.
R.
Wilburv
Laboratories.
Inc..
1995
196­
h
EC50
1422"
1T.
R.
Wilbury
Laboratories,
Inc.,
1995
Bluegill
sunfish
(Lepomis
machrochirus)
96­
h
LC50
96­
h
LC50
>
420'
569'
3M
Company,
1978
3M
Company,
1978
Water
flea
(Duphnia
magna)
48­
h
EC50
48­
h
EC50
48­
h
EC50
126*
'
3M
Environmental
Laboratory,
1982
>
1000AB
3M
Environmental
Laboratory,
1982
221B
3M
Company,
1987b
48­
h
EC50
48­
h
EC50
48­
h
EC50
292D
240
360F
EnviroSystems,
Inc.,
1990b
Ward
et
al.,
1996~
T.
R.
Wilbury
Laboratories,
Inc.,
1996b
21­
day
IC50
43'
2
1
­day
NOEC
22'
21­day
NOEC
22"

48­
h
EC50
41"
48­
h
EC50
39"
3M
Company,
1984
3M
Company,
1984
3MCompany,
1984
Ward
et
al.,
1995
Ward
et
al.,
1995
algae
(Selenastrum
capricornutum)
196­
h
EC50
1396
IWard
et
al.,
1996b
I
96­
h
EC50
96­
h
EC50
96­
h
EC50
2.1"
3.6"
1.2';
Ward
et
al.,
1995
Ward
et
al.,
1995
Ward
et
al..
1995
60
*Values
were
adjusted
to
represent
100%
active
ingredient.
*These
values
may
be
inconsistent
due
to
different
diets
tested.
"Tested
substance
was
APFO
ammonium
salt.
"Tested
substance
was
APFO
DTested
substance
was
APFO
ammonium
salt
in
50%
water.
"Tested
substance
was
APFO
ammonium
salt
in
80%
water.
FTested
substance
was
APFO
in
50%
isopropanol.
"Test
Sample:
APFO
(44%)
in
27.9%
water
and
27.2%
isopropanol
61
5.0
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