DEQ-CFW_00074839

SETAC PRESS
Environmental Toxicology and Chemistry, Vol. 22, No. 1, pp. 196-204, 2003
© 2003 SETAC
Printed in the USA
0730-7268/03 $12.00 + .00
BIOCONCENTRATION AND TISSUE DISTRIBUTION OF PERFLUORINATED ACIDS IN
RAINBOW TROUT (ONCORHYNCHUS MYKISS)
JONATHAN W. MARTIN,*t SCOTT A. MABURY,'* KEITH R. SOLOMON, f and DEREK C.G. MUIR§
tUnvversity of Guelph, Department of Environmental Biology, Bovey Building, Guelph, Ontario N1G 2W1, Canada
*University of Toronto, Department of Chemistry, 80 St. George Street, Toronto, Ontario MSS 3H6, Canada
§Environment Canada, National Water Research Institute, 867 Lakeshore Road, Burlington, Ontario L7R 4A6
(Received 30 January 2002; Accepted 5 July 2002)
Abstract —Rainbow trout (Oncorhynchus mykiss) were exposed simultaneously to a homologous series of perfluoroalkyl
carboxylates and sulfonates in a flow -through system to determine compound -specific tissue distribution and bioconcentration
parameters for perfluorinated acids (PFAs). In general, PFAs accumulated to the greatest extent in blood > kidney > liver
> gall bladder. Carboxylates and sulfonates with perfluoroalkyl chain lengths shorter than seven and six carbons, respectively,
could not be detected in most tissues and were considered to have insignificant bioconcentration factors (BCFs). For detectable
PFAs, carcass BCFs increased with increasing length of the perfluoroalkyl chain, ranging from 4.0 to 23,000, based on wet
weight concentrations. Carboxylate carcass BCFs increased by a factor of eight for each additional carbon in the perfluoroalkyl
chain between 8 and 12 carbons, but this relationship deviated from linearity for the longest PFA tested, possibly because
of decreased gill permeability. In general, half-lives (3.9-28 d) and uptake rates (0.053-1,700 L/kg/d) also increased with
increasing length of the perfluoroalkyl chain in all tissues. Sulfonates had greater BCFs, half-lives, and rates of uptake than
the corresponding carboxylate of equal perfluoroalkyl chain length, indicating that hydrophobicity, as predicted by the critical
micelle concentration, is not the only determinant of PFA bioaccumulation potential and that the acid function must be
considered.
Keywords—Bioaccumulation Fish Perfluorinated acid Perfluorooctane sulfonate Surfactants
INTRODUCTION
Perfluorinated acids (PFAs) are a class of anionic fluori-
nated surfactants characterized by a perfluoroalkyl chain and
a sulfonate or carboxylate solubilizing group. Perfluorinated
acids have been used increasingly over the past 20 years be-
cause of their temperature and chemical stability, lipophobic-
ity, and effectiveness as surfactants at low concentrations [1].
The total production of fluorinated surfactants (anionics, ca-
tionics, and neutrals) was 200 t in 1979 [1], whereas in 2000,
the total production of one PFA, perfluorooctane sulfonate
(PFOS), was nearly 3,000 t [2], Although this production vol-
ume represents only a small fraction of total surfactant man-
ufacturing (i.e., <<0.01%) [2,3], PFOS recently emerged as a
global contaminant after its detection in humans and wildlife
from various geographic locations.
By using liquid chromatography —tandem mass spectrom-
etry (LC -MS -MS) [4,5], human serum was found to contain
ng/ml concentrations of PFOS, perfluorchexane sulfonate
(PFHxS), perfluorooctanoate (PFOA), and perfluorooctane sul-
fonylamide [5]. Soon thereafter, wildlife samples collected in
various global locations also were observed to contain PFOS
[6,7], and organisms consuming fish, such as predatory birds
and mink, contained greater concentrations of PFOS than their
food sources [6,8]. The wide distribution of PFAs in the en-
vironment, despite low production volumes, is largely a result
of recalcitrance to biotic and abiotic degradation mechanisms
[9). Furthermore, atmospheric transport of PFAs may be ac-
To whom correspondence may be addressed
[e-mail]). The current address of J.W. Martin is
Department of Chemistry, University of Toronto, 80 St. George Street,
Toronto, Ontario M5S 3H6, Canada.
196
commodated through recently detected transient volatile de-
rivatives [10], and direct atmospheric emission of perfluo-
roalkyl carboxylates (C3-C14) occurs through thermolysis of
polytetrafluoroethylene [11]. Because of concerns regarding
their ubiquitous distribution in humans and wildlife, as well
as their environmental persistence, the main manufacturer of
PFAs currently is phasing out the production of long -chain
perfluorinated acids [2).
Dietary exposure of fish to PFAs does not result in bio-
magnification [12]; however, the major uptake route for wa-
terborne xenobiotics in fish is directly across the gills [13],
and direct uptake of chemicals from water (i.e., bioconcentra-
tion) is probably much more important than accumulation from
food (i.e., biomagnification) [14]. Many hydrocarbon surfac-
tants are known to bioconcentrate in fish [3,15-19], and the
European Economic Community has even adopted surface ac-
tivity as an indicator of a compound's bioconcentration po-
tential [20]. The predictably low Henry's law constant for PFAs
indicates that they will accumulate in the aquatic environment
[21,22], making fish a relevant test organism for PFA bioac-
cumulation testing. Under field conditions, fish collected
downstream of a fire -fighting foam spill contained higher con-
centrations of several PFAs than fish collected upstream [23].
We describe here the results of a tissue distribution and
bioconcentration study with rainbow trout (Oncorhynchus my -
kiss) exposed simultaneously to a suite of perfluoroalkyl car-
boxylates and sulfonates of varying fluorinated chain lengths
in a flow -through aqueous exposure. We report compound -
specific bioaccumulation parameters determined by liquid
chromatography —tandem mass spectrometry (LC -MS -MS),
and discuss their relation to physicochemical properties and
structure.
DEQ-CFW 00074839

Bioconcentration and tissue distribution of PFAs
Environ. Toxicol. Chem. 22, 2003 197
MATERIALS AND METHODS
Standards and reagents
Standards of potassium perfluorobutane sulfonate (PFBS),
potassium PFHxS (99.9%), and potassium PFOS (86.4%) were
provided by the 3M Company (St. Paul, MN, USA). Standards
of perfluoropentanoic acid (97%), perfluoroheptanoic acid
(PFHpA, 99%), PFOA (981/o), perfluorononanoic acid (PFNA,
970/6), perfluorodecanoic acid (PFDA, 981/6), perfluorounde-
canoic acid (950/6), perfluorododecanoic acid (PFDoA, 950/6),
and perfluorotetradecanoic acid (PFTA, 97%) were purchased
from Sigma -Aldrich (Oakville, ON, Canada), and perfluoroh-
exanoic acid (95%) was obtained from Oakwood Research
Chemicals (West Columbia, SC, USA). Ammonium acetate
(98%) and tetrabutylammonium hydrogensulfate were pur-
chased from Sigma -Aldrich, anhydrous sodium carbonate
(99.8%) was purchased from J.T. Baker (Phillipsburg, NJ,
USA), and methyl-tert-butyl ether (99.5%) was purchased
from EM Science (Gibbsburg, NJ, USA).
Fish rearing
Rainbow trout were purchased from Rainbow Springs (Tha-
mesford, ON, Canada) and were allowed to acclimate to lab-
oratory conditions for two weeks before exposure. Carbon -
filtered and dechlorinated water (with Na2S03) was maintained
at 12°C, and a 12-h photoperiod was used. Fish were fed daily
at a rate of 1.5% body weight per day, corrected for growth
throughout the experiment. Trout feed was purchased from
Martin Mills (Tavistock, ON, Canada).
Stock perfluorinated acid solution preparation
A stock solution containing all the test compounds was
produced by first dissolving PFAs in a small amount of meth-
anol, which subsequently was dissolved in 30 L of reverse -
osmosis laboratory water. The resulting solution was contained
in a polypropylene container and was stirred for 3 d before
the beginning of the experiment. After allowing the solution
to settle for several hours, some of the test compounds had
not completely dissolved, based on the appearance of a white
solid at the surface of the solution. This material was removed
by filtering the entire solution through glass (GF/C) microfiber
filters (Whatman, Kent, UK). The resulting solution was used
for exposure and was constantly stirred while being delivered
to exposure tanks.
Bioconcentration exposure
Juvenile rainbow trout (5-10 g) were exposed to a 1,000-
fold dilution of the stock PFA solution in a flow -through ex-
posure design for 12 d, followed by 33 d of depuration in clean
water. Two days before the initial exposure, fish were trans-
ferred to glass aquaria lined with polypropylene bags. Per -
fluorinated acids have been reported to bind to glass surfaces
[5], and the plastic bags were used as a precaution to minimize
equilibration time after introduction of the test compounds.
Initial biomass loading was 8 g/L in the exposure tank, and
2.5 g/L in the control tank. Fish growth was monitored by
weighing the total biomass every 2 to 3 d throughout the course
of the experiment.
Dilution water was gravity fed to each aquarium at 500 ml/
min, and a peristaltic pump delivered the stirred stock solution
into the dilution stream of the treatment tank at 0.5 ml/min.
At time 0 of the exposure, an initial volume of the stock
solution (45 ml) was added to the exposure aquarium (45 L)
to immediately achieve the desired exposure concentration.
Three fish from the exposure tank and one from the control
tank were sampled at each predetermined interval during the
uptake phase of the experiment (4.5, 9, 18, 36, 72, 144, and
288 h). At 288 h, the remaining fish were transferred to new
aquaria receiving clean water at 3 L/min. Extra care was taken
to reduce contamination of the depuration phase tanks by al-
lowing the fish to depurate for 5 min in a pail containing clean
dilution water, and then transferring the fish by net to the
depuration tanks. During the depuration phase, three fish from
the treatment tank and one from the control tank were sampled
at each time interval (4.5, 9, 18, 36, 72, 144, 288, 456, and
792 h).' During the uptake phase, water samples (1 L) were
collected below the surface in polypropylene bottles at 0.25,
4.5, 12, 18, 36, 72, 144, 197, 244, and 288 h. Water samples
were also collected at 48 and 96 h of the depuration phase
from both tanks to check for contamination.
Sampled fish were anesthetized with MS-222, a blood sam-
ple (50-200 µl) was drawn, and fish were subsequently killed
by a blow to the head and cervical dislocation. An incision
was made along the ventral surface from the anus to the gills,
and the entire liver was removed for analysis. The gut, con-
sisting of esophagus, stomach, pyloric ceca, spleen, and in-
testines, was removed but not analyzed. The blood, liver, and
carcass samples were analyzed separately for PFAs at each
sampling time to determine the kinetics of uptake and depu-
ration.
Tissue distribution exposure
Four immature rainbow trout (30-48 g) were exposed in a
separate tank under the same uptake conditions as bioconcen-
tration fish. On day 12, three fish were randomly sampled and
anesthetized with MS-222. Blood samples were immediately
drawn and the fish were euthanized by a blow to the head and
cervical dislocation. Fish were subsequently dissected to sep-
arate and remove the spleen, heart, liver, gall bladder, gonads,
gills (bones removed), adipose tissue (separated from pyloric
ceca), gut (including esophagus, stomach, intestine, and py-
loric ceca), and kidney. A small sample of white muscle was
also removed adjacent to the dorsal fin of each fish and sep-
arated from the skin.
Analysis of perfluorinated acids by liquid
chromatography -tandem mass spectrometry
Liver and blood samples were homogenized in 15-ml plastic
(polypropylene copolymer) centrifuge tubes containing 3 ml
of Na2CO3 (0.25 M), 1 ml of water, 1 ml of the ion -pairing
agent tetrabutylammonium hydrogensulfate (0.5 M adjusted to
pH 10) [5], and 100 µl (25 ng) of the internal standard, PFNA.
Carcass samples were first reduced to a fine powder by using
a mortar and pestle with liquid nitrogen, and subsequently were
homogenized in 50-ml plastic centrifuge tubes containing 10
to 20 ml of NaZC03. An exact quantity (2-4 g) of the ho-
mogenate was then transferred to a separate centrifuge tube
containing 1 ml of water, 1 ml of tetrabutylammonium hydro-
gensulfate, and 100 µl of PFNA.
The resulting homogenates were ektracted with 5 ml of
methyl-tert-butyl ether by shaking vigorously for 10 min, fol-
lowed by centrifugation to isolate the organic phase. The meth-
yl -tent -butyl ether supernatant was collected in a separate plas-
tic tube, and this extraction process was repeated once more,
combining the supernatants. The methyl-tert-butyl ether was
blown to dryness under high -purity nitrogen gas, and the an-
alytes were taken up in 1 to 2 ml of 50:50 (v/v) water:methanol
DEQ-CFW 00074840

198 Environ. Toxicol. Chem. 22, 2003 J.W. Martin et al.
Table 1. Test compound acronym, structure, ion transition monitored, and mean aqueous exposure concentration ± 1 relative standard deviation
(RSD)
Ion transition
Waterbome
monitored by
concentration
Test compound
Acronym
Structure
LC-MS-MS1
(µg/L) ± RSD (%)
Perfluorocarboxylates
Perfluoropentanoic acid
PFPA
CF3(CF,)3CO,H
263 —) 219
1.7 ± 11
Perfluorohexanoic acid
PFHxA
CF3(CF,),CO,H
313 269
1.7 ± 10
Perfluoroheptanoic acid
PFHpA
CF3(CF,)SCO,H
363 319
1.6 ± 12
Perflu0000ctanoic acid
PFOA
CF3(CF,)6COZH
413 —4 369
1.5 ± 13
Perfluorodecanoic acid
PFDA
CF3(CF2)gCO2H
513 —4 469
0.71 ± 24
Perfluoroundecanoic acid
PFUnA
CF3(CF3),CO,H
563 519
0.48 ± 26
Perfluorododecanoic acid
PFDoA
CF3(CF2),aCO2H
613 569
0.20 ± 30
Perfluorotetradecanoic acid
PFTA
CF3(CF2)12CO2H
713 —) 669
0.014 ± 30
Perfluorosulfonates
Perfluorobutane sulfonic acid
PFBS
CF3(CF2)3S03H
299 99
1.4 ± 10
Perfluorohexane sulfonic acid
PFHxS
CF3(CF2)SS03H
399 99
1.4 ± 11
Perfluorooctane sulfonic acid
PFOS
CF3(CF2),S03H
499 -) 99
0.35 ± 26
, LC -MS -MS = liquid chromatography —tandem mass spectrometry.
by vortexing for 30 s. The solution was then filtered through
0.2-µm nylon filters into polypropylene vials for analysis.
Direct analysis of uptake phase water samples was possible
for all PFAs except for PFDoA and PFTA, which were below
the limit of detection (-0.4 µg/L). For direct analysis, 1-L
samples were shaken vigorously and 1 ml was withdrawn from
below the surface of each with a polypropylene syringe. These
subsamples were then filtered through 0.2-µm nylon filters into
polypropylene shell vials for analysis. These samples were
analyzed without PFNA, and instead used external standards
and quantitation by standard curve analysis. To determine
PFDoA and PFTA uptake concentrations (36, 72, and 197 h),
and to check for PFA contamination of depuration water, 25
ml of each water sample was removed and filtered (0.2-µm
nylon) into a 50-ml plastic centrifuge tube containing 0.6 g of
Na2CO31 I g of tetrabutylammonium hydrogensulfate, and 100
µl of PFNA. Perfluorinated acids were extracted two times by
using 20-ml aliquots of methyl -tent -butyl ether, centrifuging,
combining the supernatants, and analyzing further in the same
manner as tissue samples.
Instrumental analysis was performed by LC -MS -MS by us-
ing previously described conditions [4,23]. Water and meth-
anol solvents (0.01. M ammonium acetate) were delivered at
a total flow rate of 300 µl/min by a Waters 600S controller
(Milford, MA, USA), and samples (25 µl) were injected with
a Waters 717 plus autosampler. Chromatography was per-
formed on a Genesis C8 column (2.1 X 50 mm, Jones Chro-
matography, Lakewood, CO, USA). Initial mobile phase con-
ditions were 90:10 (v/v) water:methanol for 30 s, followed by
a 10-min ramp to 0:100, a 3.5-min hold, and reverting to initial
conditions at 14 min. The detector was a Quattro LC (Micro -
mass, Manchester, UK) equipped with an electrospray interface
operating in negative ion mode. Data were acquired by MS -
MS by using a multiple reaction monitoring method that mon-
itored a single transition (parent --> daughter ion) for each
compound (Table 1). Desolvation temperature was 350°C, and
the source block was maintained at 150°C. Desolvation gas
flow was between 600 and 700 L/h, and the capillary voltage
always was 2.75 kV.
Quantitation was performed relative to PFNA by using a
standard curve constructed from known quantities of standards
extracted from water in the same manner as tissue samples.
All samples were blank subtracted before quantitation, and
standard injections were made every six to nine samples to
monitor sensitivity drift. Blank concentrations were absent for
most PFAs; however, PFOA and PFOS were detected on oc-
casion below 10% of sample concentrations.
Data analysis
Fish weight (FW) was best predicted by the exponential
growth model, FW = a•exp(g•t), where a is a constant, g is
the growth rate, and t is the time. All tissue concentrations
were corrected for growth dilution by determining the percent
increase in FW at each sampling interval, relative to t = 0,
by using the predicted exponential growth rate equation. The
depuration rate constants (kd) were determined by linear re-
gression after fitting the growth corrected depuration concen-
trations (Cfim,(,)) to the first -order decay model Cfish(,) =
a•exp(—kd•t), where a is a constant. Depuration half-life was
calculated by the formula ln(2)/kd.
Uptake rate constants (k,) were determined by using iter-
ative nonlinear regression (Systat®, Ver 9.0, Systat Software,
Richmond, CA, USA), by fitting the growth -corrected tissue
concentrations to the integrated form of the kinetic rate equa-
tion for constant aqueous exposure [14]
Cfl.,te = [(k)(CJ/(kd)]'[1 — exp(—kd't)] (1)
where C,,, is the average exposure water concentration and kd
is a fixed parameter. Bioconcentration factors (BCFs) were
calculated by the quotient (k lkd).
RESULTS AND DISCUSSION
Fish mortality, growth, and liver somatic index
Only one fish (2%) died during the uptake phase, and no
statistically significant difference was found in the rate of
growth for exposed fish relative to controls (p = 0.41; Table
2). The mean initial fish mass (Table 2) was greater in this
study relative to the dietary accumulation study [12], resulting
in a lower rate of growth. No statistically significant difference
was found between the liver somatic index of exposed and
control fish (p = 0.854; Table 2). Perfluorinated acids cause
hepatomegaly in rodents upon exposure via peroxisome pro-
liferation [24), but this effect was not apparent in these fish
under this exposure scenario based on liver mass.
DEQ-CFW 00074841

Bioconcentration and tissue distribution of PFAs Environ. Toxicol. Chem. 22, 2003 199
Table 2 Uptake and deputation phase duration, growth rate constant, and the associated coefficient of determination, mortality, and liver somatic
index (LSI) for exposed and control juvenile rainbow trout used in bioconcentration testing
Uptake
Deputation
Mean initial
Growth rate
Mortality LSIe
period (d)
period (d)
fish mass (g)
(10-3 g/d) (rz)
(e/O) (%)
Exposure tank 12
33
7.3
4.9 (0.50)
2 1.1 ± 0.03
Control 12
33
7.9
5.1 (0.43)
0 1.1 ± 0.07
a Average of all fish sampled throughout the uptake and deputation period in each tank, for each experiment. Linear regression revealed no
increase or decrease in LSI throughout.
Water concentrations
During the uptake period the average exposure water con-
centrations were between 0.014 and 1.7 µg/L for individual
PFAs (Table 2 and Fig. 1). The low exposure concentrations
for longer PFAs were a result of their low water solubility,
whereas more soluble PFAs were close to the nominal con-
centration (-2 µg/L). Perfluorinated acid concentrations were
relatively stable throughout the uptake phase after an initial
decrease between 0.25 and 24 h (Fig. 1). This decrease was
more pronounced for more hydrophobic compounds (i.e.,
PFOS > PFHxS; Fig. 113), suggesting that rapid uptake by
fish was influencing water concentrations to some extent de-
spite flow -through conditions. As biomass was removed
through sampling, and as uptake rates decreased, the water
101
103
104
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10'3
10-4
sJ i0-3
E
Im
p 103
Im
� 1 or,
c
0
10�2
c
pc
U 10°
Uri
10-2
IV
uptake I deputation
A
-
Blood
tee--'•
neum
�- PPOA
...........
. .. - . -fe- PFDA
�- PFUnA
-0- PFDOA
.�
_ . .... --a- PPTA ._ _..._......
■
water
+- PPOA
PFDA
♦
-■ MUM
Liver
Carcass
concentrations stabilized or increased slightly at later sampling
intervals, and approached the initial concentrations (i.e., 0.25
h). The mean water concentration was used as a constant es-
timate of C,,. in Equation 1 for determination of kn. Perfluor-
inated acids could not be detected in deputation water above
the limit of detection (-5 ng/L) in control or treated fish tanks.
Tissue distribution
In general, PFA concentrations were greatest in the blood
> kidney > liver > gall bladder, and lowest in the gonads >
adipose > muscle tissue (Fig. 2). Within the blood, the plasma
contained between 94 and 99% of total PFAs, with only a
minor fraction detectable in the cellular fraction. Perfluorinated
acids also were detectable in the gills, suggesting that this was
100
ltrt
1T2
10-3
100
10't
10'2
Uri
IW2
10-3
uptake I deputation
B
Blood
T18sUe
^e-- Pry
e PPHXs
water
•
. PPOS
•
-,- PFM
Liver
Carcass
- --- --- --- --- ---- ---- - --- --- --- --- ---- ----
Time (h)
Fig. 1. Growth -corrected uptake and deputation phase concentrations for (A) perfluoroalkyl carboxylates and (B) perfluoroalkyl sulfonates in
blood, liver, and carcass. Each tissue concentration represents the mean of three different fish (± 1 standard error). Aqueous exposure concentrations
during the uptake phase are shown in A and B for carboxylates and sulfonates, respectively. See Table 1 for definitions of test compounds.
DEQ-CFW 00074842

200 Environ. Tozicol. Chem. 22, 2003
Im
10
0.1
0.01
0.001
't3 �
N
l0
�
Y m
a
c
0
m
J.W. Martin et al.
a a
0 �?
a s
a
c.
Fig. 2. Perfluorinated acid concentrations, ± 1 standard error, in various fish tissues after a 12-d aqueous exposure. See Table I for definitions
of test compounds.
the site of uptake, depuration, or both, as has been determined
for other xenobiotics [13], including surfactants [15]. Recov-
ery of analytes from the heart and spleen was low (<10%
based on internal standard response), thus preventing quanti-
tation of PFAs in these tissues. In almost all surfactant bio-
concentration tests for which tissue -specific data have been
presented, the highest analyte (i.e., radiolabel) concentrations
typically are found in the gall bladder, probably a result of
metabolism in the liver [15]. The gall bladder concentrations
determined herein may be lower than in previous surfactant
studies [15] because the fish were fed daily, thus preventing
concentration and accumulation of the bile [25].
Unlike lipophilic chlorinated organic pollutants, PFAs did
not accumulate preferentially in adipose tissue. For example,
dioxin congeners accumulated in the ceca and depot fat of
rainbow trout to a greater extent than kidney, spleen, muscle,
or skin tissue [26]. The tissue distribution of PFDA in male
rats is similar to our results for rainbow trout, except that rat
liver contained by far the greatest concentrations, followed by
blood > kidney > heart > fat > testis > muscle [27). This
may be partially attributable to the lipophobic properties of
PFAs; however, PFAs also have a high affinity for plasma
albumin [28], and it has been hypothesized that PFAs may
bind to hepatic proteins such as fatty acid —binding proteins
[27], thus explaining the high PFA concentrations in blood
and liver, respectively. The relatively high concentrations of
PFAs in rainbow trout kidney may simply reflect the high
perfusion of blood, and do not automatically indicate that uri-
nary elimination is a significant depuration route. For example,
despite high kidney concentrations of PFDA in exposed rats,
the primary mode of depuration was feces, with only minor
excretion in urine [27].
The route by which humans are exposed to PFAs is not
well characterized, but consumption of fish meat may well be
a significant source of exposure. Assuming that muscle com-
prises 67% of a rainbow trout by weight [29], the meat could
therefore contain 22, 61, and 81% of the total body burden of
PFOS, PFHxS, and PFOA, respectively, based on the tissue
concentrations (Fig. 2).
Bioconcentration parameters
Similar to the results of a PFA dietary accumulation study
[12], only carboxylates with more than six perfluoroalkyl car-
bons, and sulfonates with more than four perfluoroalkyl car-
bons were detected in blood, liver, and carcass at all sampling
times (Fig. 1). Shorter PFAs are expected to have insignificant
bioconcentration potential, but their uptake or depuration ki-
netics could not be determined here, although the half-life of
PFBS in liver of rainbow trout was 3.3 d in a dietary accu-
mulation study [12]. Recovery of most PFAs from liver and
carcass is quantitative [12], and recovery from blood has al-
ready been demonstrated by Hansen et al. [5] for PFOS,
PFHxS, and PFOA. We made no attempt to monitor for me-
tabolites of PFAs because the existing literature suggests that
biotransformation will be insignificant or absent [27,30,311,
and we assumed that depuration was entirely a function of
elimination. Visual observation of depuration data indicated
possible biphasic depuration for most PFAs in blood, liver,
and carcass (Fig. 1); however, this could not be demonstrated
statistically because of the small sample size. Linear regression
DEQ-CFW 00074843

Bioconcentration and tissue distribution of PFAs Environ. ToxicoL Chem. 22, 2003 201
Table 3. Rate of uptake (k), depuration (kd), half-life (d), steady-state bioconcentration factor (BCF; estimated by k /kd), and the 12-d accumulation
ratio (AR), calculated at the end of the uptake period. Error represents ± 1 standard error, and a number in parentheses represents the coefficient
of determination (r2) for the corresponding regression analysis
Test k kd x 10-3 Half-life BCF
compound, (L/kg/d) (l/d) (d) (L/kg) 12-d AR
Carcass
PFOA
0.53 ± 0.041 (0.80)
130 ± 17 (0.76)
5.2 ± 0.67
4.0 ± 0.60
3.2
PFDA
29 ± 1.0 (0.95)
62 ± 8.2 (0.71)
11 ± 1.5
450 ± 62
350
PFUnA
120 ± 4.8 (0.95)
46 ± 6.5 (0.68)
15 ± 2.2
2,700 ± 400
1,400
PFDoA
700 ± 29 (0.94)
38 ± 5.5 (0.68)
18 ± 2.6
18,000 ± 2,700
6,600
PFTA
580 ± 26 (0.93)
24 ± 5.5 (0.45)
28 ± 6.4
23,000 ± 5,300
8.500
PFOS
53 ± 1.3 (0.98)
48 ± 6.5 (0.70)
15 ± 2.0
1,100 ± I50
690
PFHxS
0.62 ± 0.021 (0.96)
65 ± 6.2 (0.82)
11 ± 1.0
9.6 ± 0.99
7.6
Blood
PFOA
4.1 ± 0.31 (0.83)
150 ± 53 (0.83)
4.5 ± 1.6
27 ± 9.7
25
PFDA
160 ± 4.6 (0.97)
58 ± 7.2 (0.97)
12 ± 1.5
2,700 ± 350
1,900
PFUnA
500 ± 18 (0.96)
46 ± 5.3 (0.96)
15 ± 1.8
11,000 ± 1,400
5,500
PFDoA
1,700 ± 65 (0.96)
41 ± 4.3 (0.96)
17 ± 1.7
40,000 ± 4,500
18,000
PFTA
1,400 ± 120 (0.82)
48 ± 5.28 (0.82)
15 ± 1.6
30,000 ± 4,200
20,000
PFOS
240 ± 8.4 (0.90)
57 ± 6.7 (0.90)
12 ± 1.4
4,300 ± 570
3,100
PFHxS
5.3 ± 0.48 (0.76)
70 ± 6.5 (0.76)
10 ± 0.94
76 ± 9.7
59
Liver
PFOA
1.4 ± 0.13 (0.65)
180 ± 13 (0.91)
3.9 ± 0.28
8.0 ± 0.59
12
PFDA
74 ± 5.3 (0.84)
65 ± 8.9 (0.69)
11 ± 1.5
1,100 ± 180
1,100
PFUnA
260 ± 20 (0.84)
55 ± 7.7 (0.68)
13 ± 1.7
4,900 ± 770
3,800
PFDoA
910 ± 79 (0.78)
50 ± 6.7 (0.70)
14 ± 1.9
18,000 ± 2,900
11,000
PFTA
960 ± 98 (0.69)
31 ± 5.5 (0.60)
22 ± 3.7
30,000 ± 6,000
8,700
PFOS
260 ± 17 (0.88)
50 ± 7.4 (0.66)
14 ± 2.0
5,400 ± 860
2,900
PFHxS
5.8 ± 0.31 (0.91)
58 ± 7.0 (0.75)
12 ± 1.5
100 ± 13
54
, Refer to Table 1 for definitions of test compounds.
analysis of all growth -corrected depuration data resulted in
acceptable coefficients of determination (Table 3).
For detected PFAs, depuration rate constants generally de-
creased with increasing length of the perfluoroalkyl chain for
carboxylates and sulfonates in all tissues, ranging from 0.024
to 0.180/d, representing biological half-lives in the range of
3.9 to 28 d. For carboxylates, positive linear relationships were
observed between half-life and perfluoroalkyl chain length for
the carcass and liver, but not the blood (Fig. 3). In blood, the
35
30
a 25
20
Al
�_ 16
10
o
relationship deviated from linearity for the compound with the
longest chain length, PFTA, which was eliminated more quick-
ly than expected from extrapolation of the observations from
PFAs with shorter chain lengths. Perfluorotetradecanoic acid
probably partitioned out of the blood into the carcass, rather
than being eliminated, as has been shown for other anionic
surfactants [181. Based on the linear regression equations (Fig.
3), half-life increased by 2.8 to 3.8 d for each additional carbon
in the perfluoroalkyl chain. Similar analysis could not be ap-
• Perfluoroalkyl carboxylates
G Perfluoroalkyl sulfonatse
(A) carcass
(B)fiver
(C) blood
y=-22.4+3.81 x
rz= 0.97
y=-16.6+2.81 x
rz = 0.97
y=-16.8+3.12x
r2 = 0.98
II
i�
i
I
rr
§ I
zy
5.66+0.655x
y=-4.69+0.707x
n=.7+0.655x
o rz = 0.99
rz = 0.99
11 13 6 7 9 11 i3 6 7 9 11 13
Perfluoroalkyl Chain -Length
Fig. 3. Half-life (in 2/kd) and uptake rate (log kj relationships with perfluoroalkyl chain length of carboxylates and sulfonates for (A) carcass,
(B) liver, and (C) blood. Linear regression was applied to perfluoroalkyl carboxylate data, and the resulting equation and coefficient of determination
(r2) are shown. The dashed portion of the regression line is an extrapolation and indicates that perfluorotetradecanoic acid was not included in
the regression analysis.
DEQ-CFW 00074844

202 Environ. Toxicol. Chem, 22, 2003
plied to the sulfonates because only two were detectable; how-
ever, in all tissues, the half-lives of PFHxS and PFOS were
greater than those of carboxylates of equivalent perfluoroalkyl
chain length (Fig. 3).
The depuration half-lives determined in this study are
slightly greater than were determined in a dietary PFA ex-
posure study [12]. Assuming that the mode of exposure is not
important, we suggest that some of this variability may be
attributable to the difference in fish size. For example, at the
beginning of the depuration period in this study, the average
trout mass was 8.4 g, compared with only 4.0 g in the dietary
exposure study [12]. Hydrophobic organochlorine compounds
previously have been demonstrated to have longer half-lives
in larger fish [32,33], even when controlling for growth di-
lution. Body size also may help to explain the variability
among half-lives reported for humans (1,428 d [21), monkeys
(180 d [34]), and male rats (89 d [34]). In comparison, the
half-life in fish carcass was approximately 15 d for juvenile
trout, but the possibility of elimination through the gills cannot
be ruled out as an additional depuration mechanism for fish.
The rate of depuration for PFAs is generally more rapid
than for persistent organochlorine contaminants, including po-
lychlorinated biphenyls, toxaphene, hexachlorobenzene, mi-
rex, and chlorinated alkanes [33,351. However, the rate of dep-
uration for PFAs, with the exception of PFOA, is slower than
for any previously investigated surfactant in fish [15], and this
may be partially attributable to the lack of metabolism or bio-
transformation [27,30,31 ]. The relatively slow PFA depuration
half-lives, combined with the observations of high blood, liver,
and gall bladder concentrations, also support the theory that
PFAs enter into enterohepatic recirculation in fish, the process
whereby compounds are continuously recycled between the
blood, liver, gall bladder, and intestines, where resorption oc-
curs via the portal vein. This is further supported by the ob-
servation of high PFA assimilation efficiencies [12], which
indicate that intestinal resorption efficiency is similar to that
of endogenous bile acids. Use of the relatively low assimilation
efficiency of PFOA (59% [12]) as a measure of resorption
from the gut during enterohepatic recirculation could explain
why its half-life is shorter than for longer PFAs, which have
a higher resorption capacity [12]. The divergence of sulfonate
and carboxylate rates of depuration (Fig. 3) also could be
attributable to the different efficiency of resorption from the
gut during enterohepatic recirculation. For example, the as-
similation efficiency of PFHxS, which has only six perfluo-
roalkyl carbons, is greater than that for PFOA, which has seven
perfluoroalkyl carbons [12]. Also, short PFAs possibly are
eliminated in the urine to a greater extent than longer PFAs,
as has been shown for PFHpA in rats [24].
Uptake rate constants varied by a factor of 3,000 for dif-
ferent PFAs, and increased with increasing perfluoroalkyl
chain length (Fig. 3), ranging from 0.53 to 1,700 L/kg/d for
all tissues (Table 3). A similar trend, showing increased uptake
rates with increasing alkyl chain length, has been reported for
a series of cationic monoalkyltrimethyl ammonium surfactants
[36]. In general, these PFA uptake rates may be considered
very rapid and, to our knowledge, the carcass k reported here
for PFDoA exceeds any previously reported rate of uptake for
anionic surfactants (i.e., 642 L/kg/d reported by Tolls et al.
[3]). Given equal perfluoroalkyl chain lengths, sulfonates were
taken up at a greater rate than carboxylates, always lying above
the regression line for carboxylates (Fig. 3). For carboxylates,
12
10
rn s
o>
E e
ro
IL 4
OS 2
m
J.W. Martin et al.
PFDQA(C11)
I--�--i
PFunA (010)
* PFTA (C13)
PFDA (C9)
� +—PFCS (C3)
P �5 (C4) PF► ,r (C7)
PMA (C7)
0.0 0.1 0.2 0.3 0.4
Gills (µg g-1)
Fig. 4. Relationship between gill and blood concentrations, ± 1 stan-
dard error, in tissue distribution in fish. Each data point represents
the mean of three fish, and is labeled with the perfluorinated acid and
the length of the respective perfluoroalkyl chain in parentheses (e.g.,
C8 = eight carbons). See Table 1 for definitions of test compounds.
uptake increased in a linear manner with perfluoroalkyl chain
length until the length reached 13 carbons (i.e., PFTA).
The rate of uptake for PFTA, presumably the most hydro-
phobic PFA tested, was less than expected based on extrap-
olation from shorter PFAs in liver, blood, and carcass (Fig. 3).
Such a phenomenon is well characterized for hydrophobic
chlorinated contaminants, and can result from exclusion based
on limited diffusive mass transfer [37], molecular size [38],
or from reduced bioavailability because of binding on dis-
solved or colloidal organic matter in the water column [39,40].
Reduced bioavailability because of association with dissolved
and colloidal organic matter has been demonstrated for cationic
surfactants [15], and anionic surfactants [41]. The octanol—
water partition coefficient (Kow) for PFTA is unknown; how-
ever, it is feasible that the long perfluoroalkyl chain of PFTA
causes the molecule to act similar to superhydrophobic neutral
organochlorines (i.e., log Kow > 6), which interact with or-
ganic matter, resulting in an apparent reduction of the uptake
rate [39]. Alternatively, based on the ratio of concentrations
in the gills and blood plasma at day 12 of the tissue distribution
study, the gills appear to be less permeable to PFTA than
shorter PFAs (Fig. 4). Perfluorotetradecanoic acid was en-
riched at the gills relative to shorter perfluoroalkyl carboxyl-
ates and sulfonates, suggesting that gill permeation rates may
limit uptake of PFTA, as has been shown for the high molecular
weight cationic surfactants C16-alkylpyridinium and (C102-di-
alkyldimethylammonium [15]. A time course of the tissue dis-
tribution would allow for more conclusive arguments; how-
ever, gill permeation is known as the rate -limiting step for
uptake of other anionic surfactants [18].
Steady-state carcass BCFs ranged from 4.0 to 23,000, and
generally increased with the perfluoroalkyl chain length (Table
3 and Fig. 5). Similar trends, showing increased BCFs with
increased alkyl chain length, have been reported for anionic
linear alkylbenzene sulfonate surfactants [18] and some cat-
ionic surfactants [36]. For PFAs, blood and liver BCFs were
1 to 10 times greater than in the carcass (Table 3). Carboxylate
BCFs were greatest in the blood, whereas sulfonate BCFs were
greatest in the liver. Carcass BCFs closely approximate the
whole -body BCF; however, removal of the liver and a portion
of the blood from the intact fish resulted in a minor loss of
DEQ-CFW 00074845

Bioconcentration and tissue distribution of PFAs
LL 4
V
f>0
3
• Perfluoroalkyl carboxylates
o Perfluoroalkyl sulfonates
y = —5.73 + 0.915 x
II
7 9 11 13
Perfluoroalkyl Chain Length (carbons)
Fig. S. Carcass bioconcentration factor (BCF) relationship with per-
fluoroalkyl chain length. Linear regression was applied to perfluo-
roalkyl carboxylate data, and the resulting equation and coefficient of
determination (e) are shown. Perfluorotetradecanoic acid was not
included in the regression analysis and the dashed portion of the
regression line is an extrapolation of the regression equation.
total PFAs. The resulting carcass contained between 85% (i.e.,
PFHxS) and 96% (i.e., PFTA) of total PFAs, depending on the
test compound. Bioconcentration factors increased largely as
a result of increasing uptake rate, whereas depuration rate
varied only by a factor of eight for all tissues examined and
had less relative influence on the observed BCF trends.
Because steady state was not achieved for any PFA, a 12-
d accumulation ratio was calculated (Table 3) and compared
to the estimated steady-state BCF. The 12-d accumulation ratio
always was smaller than the BCF; however, for the PFAs with
the shortest chain lengths, the 12-d accumulation ratio ap-
proached the estimated BCF to within 80% (i.e., PFOA). The
tissue concentrations of PFAs with longer chain lengths were
far from steady state at the end of the uptake period, reaching
only 37% of the predicted steady-state value by day 12 (i.e.,
PFDoA and PFTA. The time to reach steady state is deter-
mined exclusively by the rate of depuration, and was predicted
to exceed 10, 43, and 120 d for PFOA, PFOS, and PFTA,
respectively [ 12]. The fact that steady state was not approached
should not influence our estimates of the BCF calculated by
the ratio of uptake and elimination rates. A previous compar-
ison of the kinetic and steady-state approaches to the study of
surfactant bioaccumulation indicated that both methods yield
similar results, and that the one -compartment, first -order model
was suitable [15].
Perfluoroalkyl carboxylate carcass BCFs (i.e., log BCF)
increased linearly with perfluoroalkyl chain length from PFOA
to PFDoA by a factor of eight for each additional fluorinated
carbon in the chain (Fig. 5). The trend deviated from linearity
for PFTA as a result of the decreased rate of uptake (Fig. 3A).
Given equal perfluoroalkyl chain length, sulfonates biocon-
centrated to a greater extent than carboxylates, largely as a
result of higher uptake rates, but also because of lower dep-
uration rates, and did not fit the same linear relationship for
carboxylates, always lying above the linear regression line
(Fig. 5). Furthermore, PFHxS, which has six perfluoroalkyl
carbons, had a carcass BCF of 9.6, whereas PFHpA, which
also has six perfluoroalkyl carbons, could not be detected in
most tissues despite a higher exposure concentration (Table
1). In a dietary exposure of fish to PFAs, perfluoroalkyl sul-
fonates also bioaccumulated to a greater extent than carbox-
Environ. Toxicol. Chem. 22, 2003 203
Potassium Perfluoroalkyl carboxylates
v Potassium PFOS
r
2-
log OX = 5.46 - 0.588 x
0
4
3 4 5 6 7
PFOSX* (8 mmol 0 )
P• ffuoroaikyi Chainlength
10 11
Fig. 6. Relationship determined by linear regression between the log
critical micelle concentration (log CMC) and perfluoroalkyl chain
length of potassium perfluoroalkyl carboxylate salts. The CMC for
the potassium salt of perfluorooctane sulfonate (PFOS-K*) fits the
relationship. All data have been taken from Kissa [421.
ylates of equivalent perfluoroalkyl chain length, and a similar
linear trend was observed for carboxylate bioaccumulation fac-
tors [12].
Bioconcentration potential historically has been well pre-
dicted by the Kow, a measure of a substance's hydrophobicity.
However, the Kow is a problematic parameter for surfactants
because of their tendency to_ aggregate at the interface of a
liquid —liquid system. Here we have used the perfluoroalkyl
chain length as a measure of hydrophobicity; however, Tolls
and Sijm [16] suggested that a more meaningful measure of
the relative hydrophobicity for surfactants is the critical mi-
celle concentration (CMC), the concentration at which half of
the molecules in solution are associated as micelles. Critical
micelle concentration data are limited or nonexistent for many
of the test compounds; however, the CMC of PFOS fits the
relationship determined for several carboxylates of variable
perfluoroalkyl lengths (Fig. 6) [42], indicating that the acid
function has little influence on CMC. Analysis of our data
indicates that bioaccumulation potential increases with de-
creasing CMC for carboxylates; however, the CMC alone does
not explain the disproportionate accumulation of sulfonates
relative to carboxylates. For PFAs, the acid function must be
considered for prediction of the bioconcentration and bioac-
cumulation potential [12]. For perfluoroalkyl carboxylates, the
slope of the log CMC data (Fig. 6) is remarkably similar, but
opposite in sign, to the plot of log uptake rate in carcass (Fig.
3A), indicating that increased uptake for longer PFAs within
a class may be a function of increasing hydrophobicity.
The linear CMC—BCF relationships, described by Tolls and
Sijm [16], had relatively low coefficients of determination,
and it was noted that the CMC might not be the most suitable
parameter for describing hydrophobicity of surfactants. The
suggestion was made that interfacial tension or reversed -phase
chromatographic retention time might be more suitable param-
eters. When using our chromatographic methods, PFOS and
PFNA (i.e., both of which have eight perfluoroalkyl carbons
and similar CMCs) could not be time -resolved on a reversed -
phase column, with PFNA eluting only 0.02 min after PFOS.
Additional physical and chemical properties for this series of
PFAs would be beneficial for future modeling efforts, partic-
ularly for the longer perfluoroalkyl carboxylates.
DEQ-CFW 00074846

204 Environ. Toxicol. Chem. 22, 2003
Acknowledgement -This work was supported by Health Canada and
Environment Canada (Ottawa, ON) through the Toxic Substance Re-
search Initiative, and the Canadian Network of Toxicology Centres.
We thank the 3M Company (St. Paul, MN, USA) for its generous
donation of perfluoroalkyl-sulfonate chemical standards.
REFERENCES
1. Fielding HC. 1979. Organofluorine surfactants and textile chem-
icals. In Banks RE, ed, Organofluorine Chemicals and Their
Industrial Applications. Ellis Horwood, Chichester, UK, pp 214-
234.
2. U.S. Environmental Protection Agency. 2001. Perfluorooctyl sul-
fonates; proposed significant new use rule. Fed Reg 65:62319-
62333.
3. Tolls J, Haller M, Graaf ID, Thijssen MATC, Sijm DTHM. 1997.
Bioconcentration of LAS: Experimental determination and ex-
trapolation to environmental mixtures. Environ Sci Technol 31:
3426-3431.
4. Moody CA, Kwan WC, Martin JM, Muir DCG, Mabury SA.
2001. Determination of perfluorinated surfactants in surface water
samples by two independent analytical techniques: Liquid chro-
matography/tandem mass spectrometry and 19F NMR. Anal
Chem 73:2200-2206.
5, Hansen KJ, Clemen LA, Ellefson ME, Johnson HO. 2001. Com-
pound specific quantitative characterization of organic fluoro-
chemicals in biological matrices. Environ Sci Technol 35:766-
770.
6. Giesy JP, Kannan K. 2001. Global distribution of perfluorooctane
sulfonate in wildlife. Environ Sci Technol 35:1339-1342.
7. Kannan K, Koistinen J, Beckman K, Evans T, Gorzelany JF, Han-
sen KJ, Jones PD, Helle E, Nyman M, Giesy JP. 2001. Accu-
mulation of perfluorooctane sulfonate in marine mammals. En-
viron Sci Technol 35:1593-1598.
8. Kannan K, Franson JC, Bowerman W W, Hansen KJ, Jones PD,
Giesy JP. 2001. Perfluorooctane sulfonate in fish -eating birds in-
cluding bald eagles and albatrosses. Environ Sci Technol 35:
3065-3070.
9. Key BD, Howell RD, Criddle CS. 1998. Defluorination of or-
ganofluorine sulfur compounds by Pseudomonas sp. strain D2.
Environ Sci Technol 32:2283-2287.
10. Martin JW, Muir DCG, Moody CA, Ellis DA, Kwan W, Solomon
KR, Mabury SA. 2002. Collection of airborne fluorinated organ-
ics and analysis by gas chromatography/chemical ionization mass
spectrometry. Anal Chem 74:584-590.
11, Ellis DA, Mabury SA, Martin JW, Muir DCG. 2001. Thermolysis
of fluoropolymers as a potential source of halogenated organic
acids in the environment. Nature 412:321-324.
12. Martin JW, Mabury SA, Solomon KR, Muir DCG. 2003. Dietary
accumulation of perfluorinated acids in juvenile rainbow trout
(Oncorhynchus mykiss). Environ Toxicol Chem 22:189-195.
13. Streit B. 1992. Bioaccumulation processes in ecosystems. Ex-
perientia 48:955-970.
14. Bruggeman WA, Martron LBJM, Kooiman D, Hutzinger O. 1981.
Accumulation and elimination of di-, tri-, and tetrachlorobiphen-
yls by goldfish after dietary and aqueous exposure. Chemosphere
10:811-832.
15. Tolls J, Kloepper-Sams P, Sijm THM. 1994. Surfactant biocon-
centration-A critical review. Chemosphere 29:693-717.
16. Tolls J, Sijm DTHM. 1995. A preliminary evaluation of the re-
lationship between bioconcentration and hydrophobicity for sur-
factants. Environ Toxicol Chem 14:1675-1685.
17. Tolls J, Sijm DTHM. 1999. Bioconcentration and biotransfor-
mation of the nonionic surfactant octaethylene glycol monotri-
decyl ether I4C-C13EO8. Environ Toxicol Chem 18:2689-2695.
18. Tolls J, Haller M, Seinen W, Sijm DTHM. 2000. LAS biocon-
centration: Tissue distribution and effect of hardness -Implica-
tions for processes. Environ Sci Technol 34:304-310.
19. Tolls J, Haller M, Labee E, Verweij M, Sijm DTHM. 2000. Ex-
perimental determination of bioconcentration of the nonionic sur-
factant alcohol ethoxylate. Environ Toxicol Chem 19:646-653.
20. European Economic Community. 1993. Council regulation 793/
93 of 23 March on the evaluation and control of the risks of
existing substances. Off J Eur Commun L:84/81.
21. U.S. Environmental Protection Agency. 2000. Sulfonated per-
J.W. Martin et al.
fluorochemicals in the environment: Sources, dispersion, fate and
effects. AR226-0620. 3M Company, St. Paul, MN.
22. U.S. Environmental Protection Agency. 1999. Determination of
the vapor pressure of PFOS using the spinning rotor gauge meth-
od. AR226-0048. Wildlife International, Easton, MD.
23. Moody CA, Martin JW, Kwan WC, Muir DCG, Mabury SA. 2001.
Monitoring perfluorinated surfactants in biota and surface water
samples following an accidental release of fire -fighting foam into
Etobicoke Creek. Environ Sci Technol 36:545-551.
24. Kudo N, Bandai N, Suzuki E, Katakura M, Kawashima Y. 2000.
Induction of perfluorinated fatty acids with different carbon chain
length of peroxisomal B-oxidation in the liver of rats. Chem Biol
Interact 124:119-132.
25. Eckert R, Randall D, Augustine G. 1996. Animal Physiology
Mechanisms and Adaptations, 3rd ed. W.H. Freeman, New York,
NY, USA.
26. Muir DCG, Yarechewski AL, Knoll A. 1986. Bioconcentration
and disposition of 1,3,6,8-tetrachlorodibenzo p-dioxin and oc-
tachlorodibenzo p-dioxin by rainbow trout and fathead minnows.
Environ Toxicol Chem 5:261-272.
27. Vanden Heuvel JP, Kuslikis BI, Van Rafelghem MJ, Peterson RE.
1991. Disposition of perfluorodecanoic acid in male and female
rats. Toxicol Appl Pharmacol 107:450-459.
28. Guy WS, Taves DR, Brey WS Jr. 1976.Organic fluorocompounds
in human plasma: Prevalence and characterization. In Filler R,
ed, Biochemistry Involving Carbon Fluorine Bonds. American
Chemical Society, Chicago, IL, pp 117-134.
29. Barron MG, Tarr BD, Hayton WL. 1987. Temperature -depen-
dence of cardiac output and regional blood flow in rainbow trout,
Salmo gairdneri Richardson. J Fish Biol 31:735-744.
30. Banks RE, Smart BE, Tatlow JC. 1994. Organofluorine Chem-
istry Principles and Commercial Applications. Plenum, New
York, NY, USA.
31. Goecke-Flora CM, Reo NV. 1996. Influence of carbon chain
length on the hepatic effects of perfluorinated fatty acids. A 19F
and 31P-NMR investigation. Chem Res Toxicol 9:689-695.
32. Sijm D, van der Linde A. 1995. Size -dependent bioconcentration
kinetics of hydrophobic organic chemicals in fish based on dif-
fusive mass transfer and allometric relationships. Environ Sci
Technol 29:2769-2777.
33. Fisk AT, Norstrom RJ, Cymbalisty CD, Muir DCG. 1998. Dietary
accumulation and depuration of hydrophobic organochlorines:
Bioaccumulation parameters and their relationship with the oc-
tanol/water partition coefficient. Environ Toxicol Chem 17:951-
961.
34. U.S. Environmental Protection Agency. 2000. SIDS draft initial
assessment report perfluorooctane sulfonic acid and its salts.
AR226-0978. 3M Company, St. Paul, MN.
35. Fisk AT, Cymbalisty CD; Bergman A, Muir DCG. 1996. Dietary
accumulation of Cl2- and CI 6-chlorinated alkanes by juvenile
rainbow trout (Oncorhynchus mykiss). Environ Toxicol Chem 15:
1175-1782.
36. Versteeg DJ, Shorter SJ. 1992. Effect of organic carbon on the
uptake and toxicity of quartenary ammonium compounds to the
fathead minnow, Pimephales promelas. Environ Toxicol Chem
11:571-580.
37. Gobas FAPC, Opperhuizen A, Hutzinger O. 1986. Bioconcentra-
tion of hydrophobic chemicals in fish: Relationship with mem-
brane permeation. Environ Toxicol Chem 5:637-646.
38. Opperhuizen A, van de Welde EW, Asyee GM, van der Steen
JMD. 1987. Uptake and elimination by fish of polydimethylsi-
loxanes (silicones) after dietary exposure and aqueous exposure.
Toxicol Environ Chem 13:265-285.
39. Gobas FAPC, Clark KE, Shin WY, Mackay D. 1989. Biocon-
centration of polybrominated benzenes and biphenyls and related
superhydrophobic chemicals in fish: Role of bioavailability and
elimination into the feces. Environ Toxicol Chem 8:231-245.
40. Verhaar HJM, De Jongh J, Hermens JLM. 1999. Modeling the
bioconcentration of organic compounds by fish: A novel ap-
proach. Environ Sci Technol 33:4069-4072.
41. Traina SJ, McAvoy DC, Versteeg DJ. 1996. Association of linear
alkylbenzenesulfonates with dissolved humic substances and its
effect on bioavailabilty. Environ Sci Technol 30:1300-1309.
42. Kissa E. 1994. Fluorinated Surfactants: Synthesis, Properties,
Applications. Marcel Dekker, New York, NY, USA.
DEQ-CFW 00074847