<EPA REGISTRATION DIVISION COMPANY NOTICE OF FILING FOR PESTICIDE
PETITIONS PUBLISHED IN THE FEDERAL REGISTER  (7/1/2007)>

<EPA Registration Division contact: [insert name and telephone number
with area code]>

<INSTRUCTIONS:  Please utilize this outline in preparing the pesticide
petition.  In cases where the outline element does not apply, please
insert “NA-Remove” and maintain the outline. Please do not change
the margins, font, or format in your pesticide petition. Simply replace
the instructions that appear in green, i.e., “[insert company
name],” with the information specific to your action.>

<TEMPLATE:>

<[Bayer CropScience]>

<[Insert petition number]>

	EPA has received a pesticide petition ([insert petition number]) from
[Bayer CropScience], [P.O. Box 12014, 2 T.W. Alexander Drive, Research
Triangle Park, NC 27709] proposing, pursuant to section 408(d) of the
Federal Food, Drug, and Cosmetic Act (FFDCA), 21 U.S.C. 346a(d), to
amend 40 CFR part 180. by establishing a tolerance for residues of
[prothioconazole and prothioconazole-desthio] in or on the raw
agricultural commodity [grain, cereal, group 15, except sweet corn,
sorghum, and rice] at [0.35] parts per million (ppm), [forage, cereal,
group 16, except sweet corn, sorghum, and rice] at [8.0] ppm, [stover,
cereal, group 16, except sweet corn, sorghum, and rice] at [10] ppm,
[hay, cereal, group 16, except sweet corn, sorghum, and rice ] at [7.0]
ppm, [straw, cereal, group 16, except sweet corn, sorghum, and rice] at
[5.0] ppm, [corn, sweet, forage] at [7] ppm, [corn, sweet, stover] at
[8] ppm and [corn, sweet, kernel plus cob with husks removed] at [0.02]
ppm.  EPA has determined that the petition contains data or information
regarding the elements set forth in section 408 (d)(2) of  FDDCA;
however, EPA has not fully evaluated the sufficiency of the submitted
data at this time or whether the data supports granting of the petition.
Additional data may be needed before EPA rules on the petition.

<A. Residue Chemistry>

<	1. Plant metabolism. [In plants, the metabolism of prothioconazole is
adequately understood for purposes of establishing these proposed
tolerances. Prothioconazole was extensively metabolized in plants with
the major residue found in all crops (wheat, peanuts and sugar beets)
being JAU6476-desthio with smaller amounts of various isomers of
JAU6476-hydroxy-desthio, and their conjugates also being found.
Triazolylalanine, triazolylhydroxypropionic acid, and triazolylacetic
acid, metabolites common to the triazole-derivative class of fungicides,
were also found. Based on the above data the residues of concern in
plants are prothioconazole and its metabolite JAU6476-desthio. The  risk
assessment for the common metabolites arising from the
triazole-derivative fungicides is being addressed by the US Triazole
Task Force.  In large animals, the nature of residues is also adequately
understood for purposes of establishing the proposed tolerances. The
residues of concern in edible tissues and milk are prothioconazole, its
metabolites JAU6476-desthio, JAU6476-4-hydroxy, and their conjugates
that can be converted to these three compounds by acid hydrolysis.]>

<	2. Analytical method. [The analytical method for determining residues
of concern in plants extracts residues of prothioconazole and
JAU6476-desthio and converts the prothioconazole to JAU6476-desthio and
JAU6476-sulfonic acid. Following addition of internal standards the
sample extracts are analyzed by LC/MS/MS. Radiovalidation and
independent laboratory validation have shown that the method adequately
quantifies prothioconazole residues in treated commodities. The
analytical method for analysis of large animal tissues includes
extraction of the residues of concern, followed by addition of an
internal standard to the extract. The extract is then hydrolyzed to
release conjugates, partitioned and analyzed by LC/MS/MS as
prothioconazole, JAU6476-desthio and JAU6476-4-hydroxy. The method for
analysis of milk eliminated the initial extraction step in the tissue
method.]>

	3. Magnitude of residues. [Residue data on wheat and barley have
previously been submitted to the Agency under PP 4F6830. These data were
used to establish tolerances wheat and barley as published in 72 FR
11777. In addition to data on wheat and barley, additional magnitude of
residue data from foliar applications of prothioconazole on corn (field,
pop and sweet) are being submitted to support this petition. A total of
12 sweet corn field trials, 20 field corn field trials and three pop
corn field trials were conducted to measure the magnitude of
prothioconazole residues in/on corn commodities following four
applications of JAU 6476 480 SC formulation at a rate of 200 g ai/ha
with an application interval of 5-7 days. Data submitted with this
petition support the proposed tolerances for field corn, sweet corn and
pop corn commodities noted above.

Processing data other than that submitted to the Agency under PP 4F6830
are being submitted on corn grain and corn processed commodities
following exaggerated rate applications of prothioconazole to field
corn. These data show no concentration in any of the required processed
commodities. Total prothioconazole residue did concentrate in corn
aspirated grain fractions. However, the resulting residue would be less
than that already establish for wheat as published in 72 FR 11777.]

<B. Toxicological Profile>

<	1. Acute toxicity.  [Prothioconazole exhibits very low acute oral,
dermal and inhalation toxicity. The acute oral LD50 for prothioconazole
in rats was >6200 mg/kg/day while the dermal LD50 was >2000 mg/kg/day
and the four-hour rat inhalation LC50 >4990 mg/m3. Prothioconazole was
not irritating to the eye or skin of rabbits and did not cause skin
sensitization in guinea pigs.]>

<	2. Genotoxicty. [Prothioconazole is not mutagenic. Slightly positive
and equivocal responses seen in in vitro cytogenetic and UDS assays were
negative in in vivo tests.]>

<	3. Reproductive and developmental toxicity. [Prothioconazole is not
considered a primary reproductive toxicant in a two-Generation rat
reproduction study. The NOAEL for parental toxicity is 9.7 mg/kg bw/day
and the NOAEL for reproductive and neonatal effects is 95.6 mg/kg
bw/day. In the rat oral developmental toxicity study the fetal NOAEL was
500 mg/kg bw/day and the maternal NOAEL was 80 mg/kg bw/day. A dose
level of 1000 mg/kg bw/day administered to the rat by the dermal route
does not elicit developmental toxicity. In the rabbit, prothioconazole
does not produce developmental toxicity at dose levels exceeding the
maximum tolerated dose. Thus, prothioconazole is not teratogenic in
either the rat or rabbit.]>

<	4. Subchronic toxicity. [90-day feeding studies conducted in rats,
mice, and dogs showed the liver and/or kidney to be the target organs.
In these subchronic studies NOAELs were established at 100 mg/kg/day for
the rat and 25 mg/kg/day for the mouse and dog.]>

<	5. Chronic toxicity. [Prothioconazole is not carcinogenic in either
the rat or the mouse even at dose levels in the rat exceeding the
maximum tolerated dose. The liver and kidney are target organs of
prothioconazole for non-neoplastic toxicity in the rat, mouse and dog.
The lowest NOAEL established on the basis of long-term toxicity studies
is 5 mg/kg bw/day in the rat chronic toxicity/carcinogenicity study and
the dog 52-week toxicity study.]>

<	6. Animal metabolism. [Metabolism and pharmacokinetic studies in the
rat demonstrate that prothioconazole is rapidly absorbed, metabolized
and eliminated. The primary metabolites included JAU6476-desthio, as
well as S- and O-glucuronides of prothioconazole.]>

<	7. Metabolite toxicology. [The major metabolite of prothioconazole in
the rat, plants and large animals is JAU6476-desthio. Therefore, a
comprehensive evaluation of the mammalian toxicity of JAU6476-desthio
has been performed.

The acute oral, dermal, and inhalation toxicity of JAU6476-desthio in
the rat is low (LD50 values: 2806/2506mg/kg [male/female], >5000mg/kg,
4-hour LC50: >5077mg/m3, respectively). It is non-irritating to skin and
only very slightly irritating to eyes and has no cutaneous sensitizing
potential by topical application.

The short and long-term oral toxicity of JAU6476-desthio has been
investigated by dietary administration in the rat, mouse and dog. The
common target organ in all 3 species is the liver. Secondary effects on
the thyroid in the rat and dog were associated with increased hepatic
enzyme induction. In the long term studies, 52 weeks in the rat and 30
weeks in the dog, the NOAEL for the rat and the dog were 1.1 and 10.1
mg/kg bw/day, respectively. JAU6476-desthio is not carcinogenic in
either the rat or the mouse. The NOAEL selected for human risk
assessment is 1.1 mg/kg bw/day established in the combined chronic
toxicity and carcinogenicity study in the rat.

JAU6476-desthio was negative for mutagenicity and genotoxicity in all in
vitro and in vivo studies both with and without metabolic activation.

In a two-generation reproduction study the maternal NOAEL was 2.7 mg/kg
bw/day and the NOAEL for reproductive and neonatal effects 10.0 mg/kg
bw/day. The main targets at the highest dose were dystocia (probably
secondary to hepatic toxicity), decreased pup viability, growth
retardation and low incidence of cleft palates.

In the developmental toxicity studies cleft palates were seen in the rat
and rabbit at the highest dose levels and supernumerary ribs in the rat
at lower dose levels. The oral NOAEL is 1 mg/kg bw/day in the rat and 2
mg/kg/day in the rabbit.

A developmental neurotoxicity study (DNT) in rats was conducted with
JAU6476-desthio as the anticipated dietary exposure is mainly to
JAU6476-desthio and not the parent. No evidence of neurotoxicity was
seen in this study, including no compound-related effects involving
neurobehavioral tests and neuropathology.  Compound-related effects were
limited to overt toxicity and developmental effects which were
consistent with effects seen in other developmental toxicity studies.
The overall NOAEL was 3.6 mg/kg bw/day.]>

<	8. Endocrine disruption. [EPA is required under the FFDCA, as amended
by FQPA, to develop a screening program to determine whether certain
substances (including all pesticide active and other ingredients) “may
have an effect in humans that is similar to an effect produced by a
naturally occurring estrogen, or other such endocrine effects as the
Administrator may designate.”  Following recommendations of its
Endocrine Disruptor Screening and Testing Advisory Committee (EDSTAC),
EPA determined that there was a scientific basis for including, as part
of the program, the androgen and thyroid hormone systems, in addition to
the estrogen hormone system.  EPA also adopted EDSTAC’s recommendation
that the Program include evaluations of potential effects in wildlife.
For pesticide chemicals, EPA will use FIFRA and, to the extent that
effects in wildlife may help determine whether a substance may have an
effect in humans, FFDCA authority to require the wildlife evaluations.
As the science develops and resources allow, screening of additional
hormone systems may be added to the Endocrine Disruptor Screening
Program (EDSP).

When additional appropriate screening and/or testing protocols being
considered under the Agency’s EDSP have been developed, clothianidin
may be subjected to further screening and/or testing to better
characterize effects related to endocrine disruption.]>

<C. Aggregate Exposure>

<	1. Dietary exposure. [An aggregate risk assessment was conducted for
all registered and pending uses of prothioconazole including the
proposed uses on field, sweet and pop corn. The acute toxicity endpoint
used in the acute assessment was the NOAEL of 2.0 mg/kg/day from the
prothioconazole-desthio rabbit developmental toxicity study as described
in the Agency’s HED risk assessment memorandum on prothioconazole
dated October 13, 2006. Application of a 1000x uncertainty factor
resulted in an acute reference dose (aRfD) of 0.002 mg/kg/day. The
chronic reference dose (cRfD) of 0.001 mg/kg/day resulted from
application of a 1000x uncertainty factor to the NOAEL of 1.1 mg/kg/day
from the rat chronic\oncogenicity prothioconazole-desthio toxicity
study.

Results from the acute and chronic dietary exposure assessment described
below demonstrate a reasonable certainty that no harm to the overall
U.S. population or any population subgroup will result from the use of
prothioconazole on the above crops.]>

<	i. Food. [An acute, Tier 2 dietary (food and drinking water) risk
assessment was conducted. For the population subgroup females 13-49
years old, the acute analysis resulted in an exposure of 73% of the
aRfD. Results from a chronic, Tier 2 dietary (food and water) risk
assessment indicated that the most highly exposed population subgroup
was all infants with an exposure equal to 93% of the cRfD. Chronic
exposure to the overall U.S. population was 34% of the cRfD.]>

<	ii. Drinking water. [No monitoring data are available for residues of
prothioconazole in drinking water, and EPA has established no health
advisory levels or maximum contaminant levels for residues of
prothioconazole in drinking water. In order to address drinking water,
water was included in the dietary analysis for food given above.]>

<	2. Non-dietary exposure. [Prothioconazole is not registered for
residential uses nor are any registrations pending for such uses.]>

<D. Cumulative Effects>

<	[Prothioconazole is a member of the conazole family of fungicides. The
cumulative effects of the primary common metabolites [1,2,4-triazole
(1,2,4-T) and the two conjugates (TA and TAA)] were addressed in two
assessments  by EPA (one for 1,2,4-T and one for TA and TAA). Both
assessments showed acceptable dietary exposure for all of the
metabolites. ]>

<E. Safety Determination>

<	1. U.S. population. [In the Agency’s HED risk assessment memorandum
on prothioconazole dated October 13, 2006, it is stated that
quantification of acute risk to the general population including infants
and children is not required. Based on the conservative exposure
assumptions described above and on the completeness of the toxicity
data, it can be concluded that total food and drinking water exposure to
prothioconazole from all proposed uses will utilize 73% of the acute RfD
for females 13-49 years old and 34% of the chronic RfDs for the overall
U.S. population. EPA generally has no concerns for exposures below 100%
of the RfD, because the RfD represents the level at or below which daily
aggregate exposure over a lifetime will not pose appreciable risks to
human health.]>

<	2. Infants and children. [In the Agency’s HED risk assessment
memorandum on prothioconazole dated October 13, 2006, it is stated that
quantification of acute risk to the general population including infants
and children is not required. For acute exposure, the Agency used the
population subgroup females 13-49 years old. In Bayer’s acute
assessment the exposure for females 13-49 years old equaled 73% of the
acute RfD. For chronic exposure the most highly exposed population
subgroups (all infants), utilized 93% of the chronic RfD. EPA generally
has no concerns for exposures below 100% of the RfD, because the RfD
represents the level at or below which daily aggregate exposure over a
lifetime will not pose appreciable risks to human health.]>

<F. International Tolerances>

<	[There are no Codex MRLs currently established for prothioconazole.]>

 PAGE

 PAGE   6


EPA REGISTRATION DIVISION COMPANY NOTICE OF FILING FOR PESTICIDE PETITIONS PUBLISHED IN THE FEDERAL REGISTER

EPA Registration Division contact: [Mark Dow<PHONE_NUMBER>]

INSTRUCTIONS:  Please utilize this outline in preparing the pesticide petition.  In cases where the outline element does not apply, please insert "NA-Remove" and maintain the outline. Please do not change the margins, font, or format in your pesticide petition. Simply replace the instructions that appear in green, i.e., "[insert company name]," with the information specific to your action.

TEMPLATE:

Pennzoil-Quaker State Company

[2E8049]

	EPA has received a pesticide petition ([2E8049]) from Pennzoil-Quaker State Company, 700 Milam Street, Houston, TX 77002 c/o Wagner Regulatory Associates, 7217 Lancaster Pike, Suite A, Hockessin, DE 19707  proposing, pursuant to section 408(d) of the Federal Food, Drug, and Cosmetic Act (FFDCA), 21 U.S.C. 346a(d), to amend 40 CFR part 180.910 to establish an exemption from the requirement of a tolerance for Distillates (Fishcher-Tropsch), heavy, C18-C50, branched, cyclic and linear, CAS No. 848301-69-9 when used as an inert ingredient in pesticide formulations as a solvent, diluent and dust suppressant without limitations in pesticide formulations.  EPA has determined that the petition contains data or information regarding the elements set forth in section 408 (d)(2) of  FFDCA; however, EPA has not fully evaluated the sufficiency of the submitted data at this time or whether the data supports granting of the petition. Additional data may be needed before EPA rules on the petition.

A. Residue Chemistry

Because an exemption from the requirement of a tolerance is requested, no field residue studies are required.

	1. Plant metabolism. "NA-Remove"

	2. Analytical method. "NA-Remove"   	3. Magnitude of residues. "NA-Remove"   B. Toxicological Profile

	1. Acute toxicity.  This substance was tested in an acute oral toxicity study according to the fixed dose method (OECD 420). Following a preliminary test in which there were no deaths at a dose level of 5000 mg/kg, an additional four fasted Sprague-Dawley DC strain rats received a dose of 5000 mg of the undiluted test material per kg body weight by gavage. There were no signs of toxicity, all animals showed normal gains in body weight, and there were no abnormalities on necropsy. The acute oral LD50 was therefore >5000 mg/kg.  In a dermal irritation study with rabbits (OECD 404), this substance was non-irritating at all time points of 24, 48 and 72 hours following application.  In an eye irritation study with rabbits (OECD 405), this substance was non-irritating at all time points of 24, 48 and 72 hours following application.  In a skin sensitization study in guinea pig (OECD 406), this substance was found not sensitizing to the test animals skin.

	2. Genotoxicty. This substance has been tested in a bacterial mutagenicity study (OECD 471) using Salmonella typhimurium strains TA 98, TA 100, TA 1535, TA 1537 and Escherichia coli WP2 uvrA.. The test material was dissolved in tetrahydrofuran at concentrations up to 5000 ug/plate. Appropriate solvent and positive controls were included and gave expected results. No toxicity to bacterial cells was observed. No significant increase in the number of revertants was observed at any concentration with and without metabolic activation in any of the strains tested. The results were confirmed in a repeat experiment; both experiments used the direct plate incorporation method.

An in vitro micronucleus study has been conducted (OECD draft guideline 487). No increase in the incidence of micronuclei was observed in duplicate cultures of human lymphocytes at any concentration in either the initial experiment (4 hour exposure, 16 hour expression, with and without metabolic activation) or the repeat experiment (20 hour exposure without metabolic activation; 4 hour exposure, 16 hour expression, with metabolic activation). No test material induced toxicity was observed.

Further evidence of the lack of effects on chromosomes in vitro was obtained when tested according to OECD 473. No statistically significant increase in the frequency of cells with chromosome aberrations was observed in either the initial or the repeat experiment when tested with and without metabolic activation up to a dose level that was limited by the onset of precipitate. Appropriate solvent and positive controls were included and gave expected results.

	3. Reproductive and developmental toxicity.    In a two-generation reproductive toxicity study, this substance [0 (vehicle), 50, 250, or 1000 mg/kg/day)] was given orally to F0 and F1 parental male and female rats for at least 70 days prior to mating and throughout the 14-day mating, post mating holding (for males), and gestation and lactation (for females) periods. There was a test item-related, non-adverse decrease in anogenital distance in F1 females born from dams given 250 and 1000 mg/kg/day. Test item-related histopathological lesions were identified in the lungs (chronic interstitial/alveolus inflammation) of both males and females of the F0 and F1 generations with corresponding macroscopic findings and/or increased lung weights and the kidneys (renal tubule hyaline droplets likely representing α2μ-globulin) of the F1 males only. The lung lesions were most likely secondary to aspiration of the test material and therefore not relevant for human risk assessment. The renal effects are a well-known male rat specific effect which is induced by hydrocarbons and has no relevance for humans. An equivocal, non-adverse slight decrease in F2 pup brain weights was noted. Based on the absence of adverse, test item-related findings on the integrity and performance of the male and female reproductive systems, the absence of adverse findings directly attributable to the test item in non-reproductive tissues, and the absence of adverse effects on in-life parameters (such as body weight, feed consumption, and clinical observations), the dosage level of 1000 mg/kg/day was considered to be the no-observed-adverse-effect level (NOAEL) for reproductive and systemic toxicity.

Administration of this substance in a two- generation reproductive toxicity study in male and female rats at dosages up to 1000 mg/kg/day had no effect on reproductive performance or gestation length and parturition of both the F0 and F1 parental generations.  Andrology parameters for F0 and F1 males were unaffected by test item administration. In addition, there were no test item-related effects on F1 and F2 litter parameters, postnatal survival, physical condition/mortality, and pup body weights. Vaginal patency of F1 females and balanopreputial separation in F1 males were also unaffected by test item administration. Based on the absence of adverse test item related findings, the top dose of 1000 mg/kg/day was considered to be a NOAEL for developmental toxicity.

	4. Subchronic toxicity. The oral administration of this substance to rats for a period of ninety consecutive days at dose levels of 50, 200 and 1000 mg/kg/day in an OECD 408 study resulted in treatment-related effects in animals of either sex treated with 1000 and 200 mg/kg/day. The "no observed effect level" (NOEL) was considered to be 50 mg/kg/day. However the effects detected at 200 and 1000 mg/kg/day in the lungs and mesenteric lymph nodes were considered to be secondary to aspiration following the oral gavage and a normal response of the lymph nodes clearing the material, respectively, and were therefore not considered to be an adverse effect of treatment.  The "No Observed Adverse Effect Level" (NOAEL) was therefore considered to be 1000 mg/kg/day.

Distillates (Fishcher-Tropsch), heavy, C18-C50, branched, cyclic and linear are made up of linear, cyclic and branched alkanes.  Animal studies on a variety of linear, branched and cyclic alkanes have shown that these substances are not neurotoxic with the exception of n-hexane due to its metabolism to 2,5-hexadione.  However, distillates (Fishcher-Tropsch), heavy, C18-C50, branched, cyclic and linear do not contain n-hexane and cannot be metabolized to its neurotoxic metabolites.  Furthermore, acute toxicity tests with very high doses (i.e. 5000 mg/kg body weight in the rat) of distillates (Fishcher-Tropsch), heavy, C18-C50, branched, cyclic and linear did not produce any neurotoxicity.  Also neurotoxic effects were not observed in 28-day and 90-day studies.

	5. Chronic toxicity. There is no evidence to suggest that linear, branched and cyclic alkanes present in this substance are direct-acting carcinogens. No evidence of hyperplasia or any other histopathological indications for effects related to carcinogenicity were seen in any of the available repeated-dose toxicity studies conducted for this substance and the available data indicates that this substance is not genotoxic.

	6. Animal metabolism. Due to their complex, multi-constituent nature, there are no substance-specific absorption, metabolism, distribution and excretion studies done specifically with distillates (Fishcher-Tropsch), heavy, C18-C50, branched, cyclic and linear. However, studies with various complex mixtures of normal, branched and cyclic petroleum derived hydrocarbons show that absorption is inversely correlated with carbon chain length and independent of isomeric form, preparation process, or type of product. Constituents with carbon chain lengths up to approximately C30-C35 are absorbed through the gastro-intestinal tract for 2-20%.  However, when present as mineral oil hydrocarbons of higher molecular weight, corresponding to approximate chain lengths greater than C32-C35 are not absorbed to any significant extent.  For distillates (Fishcher-Tropsch), heavy, C18-C50, branched, cyclic and linear, the majority of the constituents are of carbon chain lengths ranging from C27-C50 and are likely to be excreted, unabsorbed, in the feces.

	7. Metabolite toxicology. [No data is available]

	8. Endocrine disruption. Administration of this substance in a two- generation reproductive toxicity study in male and female rats at dosages up to 1000 mg/kg/day had no effect on the integrity and performance of reproductive systems or gestation length and parturition of both the F0 and F1 parental generations.  Andrology parameters for F0 and F1 males were unaffected by test item administration. In addition, there were no test item-related effects on F1 and F2 litter parameters, postnatal survival, physical condition/mortality, and pup body weights. Vaginal patency of F1 females and balanopreputial separation in F1 males were also unaffected by test item administration. Based on the absence of adverse test item related findings, the top dose of 1000 mg/kg/day was considered to be a NOAEL for reproductive and developmental toxicity. Therefore within the parameters of this study no harmful effects were observed on reproduction, development or overall health resulting from the interaction of distillates (Fishcher-Tropsch), heavy, C18-C50, branched, cyclic and linear with the test animal hormone system.

C. Aggregate Exposure

	1. Dietary exposure. Dietary exposure has been assessed using the EPA I-DEEM model (Inert Dietary Exposure Evaluation Model).

	i. Food. The I-Dietary Exposure Evaluation Model (DEEM) is a highly conservative model with the assumption that the residue level of the inert ingredient would be no higher than the highest tolerance for a given commodity.  Implicit in this assumption is that there would be similar rates of degradation between the active and inert ingredient (if any) and that the concentration of inert ingredient in the scenarios leading to these highest of tolerances would be no higher than the concentration of the active ingredient. The model assumes 100 percent crop treated (PCT) for all crops (every food eaten by a person each day has tolerance level residues.)

	ii. Drinking water. For the purpose of the screening level dietary risk assessment to support this request for an exemption from the requirement of a tolerance for distillates (Fishcher-Tropsch), heavy, C18-C50, branched, cyclic and linear, a conservative drinking water concentration value of 100 parts per billion (ppb) based on screening level modeling was used to assess the contribution to drinking water for the chronic dietary risk assessments for parent compound. These values were directly entered into the dietary exposure model.

	2. Non-dietary exposure. Distillates (Fishcher-Tropsch), heavy, C18-C50, branched, cyclic and linear are not currently used as inert ingredients in pesticide products that are registered for any use patterns that involve residential uses nor are there any other non-pesticidal residential uses for these inert ingredients, thus no residential exposures are expected.

D. Cumulative Effects

	No information is available to indicate that Distillates (Fishcher-Tropsch), heavy, C18-C50 branched, cyclic and linear share a common mechanism of toxicity with any other substances and that it does not appear to produce a toxic metabolite produced by other substances.  For the purposes of this tolerance action, it is assumed that Distillates (Fishcher-Tropsch), heavy, C18-C50, branched, cyclic and linear does not have a common mechanism of toxicity with other substances.

E. Safety Determination

In a 90-day repeated dose oral (gavage) toxicity study in the rat (OECD Guideline 408 and EPA 870.3100), animals were administered doses of the substance at 0, 50, 200, and 1000 mg/kg/day. Treatment-related effects in the lungs and mesenteric lymph nodes were seen at 200 and 1000 mg/kg/day dose levels, but the effects were considered to be adaptive and not to be an adverse effect of treatment. As such, the no observable effect level (NOEL) was considered to be 50 mg/kg/day, but the no observable adverse effect level (NOAEL) was considered to be 1000 mg/kg/day.

In a 2-generation reproduction study in rats dosed by oral gavage at 0, 50, 250, and 1000 mg/kg/day, the systemic NOAEL was 1000 mg/kg/day.  Therefore, the overall NOAEL for repeat dose exposures to distillates (Fishcher-Tropsch), heavy, C18-C50, branched, cyclic and linear is taken to be 1000 mg/kg/day.  Dividing the NOAEL by an uncertainty factor of 100 to account for inter- and intra-species variability, the chronic reference dose (cRfD) is therefore 10 mg/kg/day.

Finally, the chronic population adjusted dose is calculated by dividing the cRfD by the FQPA factor, which is assumed to be 1X for Distillates (Fishcher-Tropsch), heavy, C18-C50, branched, cyclic and linear.  An additional FQPA factor (10X) is not required because effects do not tend to become more severe with higher doses and an adequate and reliable toxicity database is available.

The dietary exposures (food plus drinking water) for Distillates (Fishcher-Tropsch), heavy, C18-C50, branched, cyclic and linear are expressed as a percentage of the proposed aPAD and cPAD for the general U.S. population and children 1-2 years old, the most highly exposure subpopulation.

	1. U.S. population. Based on I-DEEM screening level calculations, chronic dietary exposures are estimated to be 1.9% of the cPAD for the general U.S. population and acute dietary exposures are estimated to be 1.1% of the aPAD for the general U.S. population. These estimates demonstrate that the I-DEEM screening-level chronic and acute dietary exposure estimates (food plus drinking water) are below the cPAD and aPAD and are therefore below the level of concern for the general U.S. population. The screening-level exposure calculations reflect the default assumption of up to 50% inert ingredient in the pesticide product and, as such, reflect no concentration limit on distillates (Fishcher-Tropsch), heavy, C18-C50, branched, cyclic and linear in formulated pesticide products.

	2. Infants and children. Based on I-DEEM screening level calculations, chronic dietary exposures are estimated to be 6.2% of the cPAD for children 1-2 years old and acute dietary exposures are estimated to be 3.3% of the aPAD for children 1-2 years old.  These estimates demonstrate that the I-DEEM screening-level chronic and acute dietary exposure estimates (food plus drinking water) are below the cPAD and aPAD and are therefore below the level of concern for infants and children.

F. International Tolerances

	No international tolerances have been established for Distillates (Fishcher-Tropsch), heavy, C18-C50, branched, cyclic and linear.


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

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

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

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

Statistically significant increases ( p< 0.01) in the number of pups found dead were observed on lactation day 1 in the 3 and 10 mg/ kg/ day groups. According to the study authors, this was not considered to be treatment related because they did not occur in a dose related manner and did not appear to affect any other measures of pup viability including numbers of surviving pups per litter and live litter size at weighing. An independent statistical analysis was conducted by US 39 EPA ( 2002b). No significant differences were observed between dose groups and the response did not have any trend in dose.

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

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

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

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

Therefore, under the conditions of the study, the LOAEL for F0 parental males is considered to be 1 mg/ kg/ day, the lowest dose tested, based on significant increases in the liver and kidney weights to terminal body weight and to brain weight ratios. A NOAEL for the F0 parental males could not be determined since treatment related effects were seen at all doses tested. 40 The NOAEL and LOAEL for F0 parental females are considered to be 10 and 30 mg/ kg/ day, respectively, based on significant reductions in kidney weight and kidney weight to terminal body weight and to brain weight ratios observed at the highest dose.

The LOAEL for F1 generation males is considered to be 1 mg/ kg/ day, based on significant, dose related decreases in body weights and body weight gains ( observed prior to and during cohabitation and during the entire dosing period), and in terminal body weights; and significant changes in absolute liver and spleen weights and in the ratios of liver, kidney, and spleen weights to brain weights. A NOAEL for the F1 males could not be determined since treatmentrelated effects were seen at all doses tested.

The NOAEL and LOAEL for F1 generation females are considered to be 10 and 30 mg/ kg/ day, respectively, based on statistically significant increases in postweaning mortality, delays in sexual maturation ( time to vaginal patency), decreases in body weight and body weight gains, and decreases in absolute food consumption, all observed at the highest dose tested.

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

It is important to note that the LOAELs and NOAELs summarized above are for the entire duration of the study, and therefore represent effects resulting from developmental and/ or adult exposures. These effect levels differ from those described in section 5.1. The LOAELs and NOAELs described in section 5.1 only refer to effects resulting from developmental exposures.

4.0 Exposure Characterization PFOA has been detected in human serum of workers occupationally exposed to APFO, and it has also been measured in the general population. In general the levels in the general population are much lower than in the workers. However, it should be noted that the highest levels reported to date in the general population are similar to some of the lowest levels in workers exposed to PFOA occupationally. It is not known what the environmental concentrations of APFO are or the pathways of exposure to the general population.

4.1 Occupational exposures 3M has been offering voluntary medical surveillance to workers at plants that produce or use perfluorinated compounds since 1976. Serum PFOA levels have been measured and reported since 1993. Prior to this time, only total organic fluorine was measured. The results of biomonitoring for PFOA have been reported for 3 plants: Cottage Grove, MN; Decatur, AL and Antwerp, Belgium. Surveillance years include 1993, 1995, 1997, 1998, and 2000, although not all of the plants offered surveillance in all of these years. The 1998 data reported for the Decatur plant consist of a random sample of employees; however, volunteers participated in all of the 41 other sampling periods for all of the plants. The results of these studies are summarized in Table 2.

Table 2. Summary of Occupational Exposures ( ppm)

Plant Arithmetic Mean Range Geometric Mean 95% CI Cottage Grove 1997 ( n = 7 4) 1995 ( n = 80) 1993 ( n = 111) 6.4 6.8 5.0 0.1  81.3 0  114.1 0  80.0 NA NA NA NA NA NA Decatur 2000 ( n = 263) 1998 ( n = 126) 1997 ( n = 126) 1995 ( n = 90) 1.78 1.54 1.57 1.46 0.04  12.7 0.02  6.76 NA NA 1.13 0.9 NA NA 0.99  1.3 0.72  1.12 NA NA Antwerp 2000 ( n = 258) 1995 ( n = 93) 0.84 1.13 0.01  7.04 0  13.2 0.33 NA 0.27  0.4 NA Building 236 2000 ( n = 45) 0.106 0.008  0.668 0.053 0.037  0.076 Mean serum PFOA levels have increased slightly at both the Cottage Grove and Decatur plants since 1993. Workers at the Cottage Grove plant, where PFOA exposures are highest, have the highest PFOA serum levels. The latest sample was in 1997 ( Olsen, et al., 1998b). The mean serum PFOA level was 6.4 ppm ( range = 0.1   81.3 ppm). Only 74 employees participated in the 1997 surveillance. The eligible voluntary participation rates ranged from approximately 50% in 1997 to 70% in 1993.

At the Decatur plant, 263 of 500 employees participated in 2000 ( Olsen, et al., 2001a). The mean serum PFOA level was 1.78 ppm. It was higher in males ( n = 215) than females ( n = 48), 1.90 and 1.23 ppm, respectively. In addition, male production employees had higher mean serum levels ( 2.34 ppm). Five employees had serum levels greater than 5 ppm, the Biological Limit Value established by the 3M Exposure Guideline Committee. Cell operators had the largest increase in serum PFOA between 1998 and 2000. The highest level was in a chemical operator on the Scotchgard team ( 12.70 ppm). The mean level for the rest of the members of the team was 5.06 ppm ( range 5  9 ppm). Other job categories did not exhibit such a large increase. 3M reports that this is due to increased PFOA production at the Decatur plant beginning in 1999. Serum PFOA levels for the Antwerp plant are lower than at Decatur or Cottage Grove, and have decreased slightly since 1995 ( Olsen, et al., 2001b). Participation in medical surveillance at the 42 Antwerp plant was the highest it had ever been in 2000 ( 258 volunteers out of 340 workers). The mean serum PFOA level was 0.84 ppm, and the highest serum level reported was 7.04 ppm. As in the Decatur plant, males ( n = 209) had higher mean serum PFOA levels ( 1.03 ppm) than females ( n = 49, 0.07 ppm). Three employees had levels greater than 5 ppm.

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

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

4.3 General Population Exposures Data on PFOA levels in the general population are very limited. They are very recent so that trends over time cannot be established. The mean serum PFOA levels are lower in the general population than in workers exposed to PFOA. The available data are summarized in Table 3. 43 Table 3. Summary of General Population Exposures ( ppb)

Sample Arithmetic Mean Range Geometric Mean 95% CI Pooled Samples Commerical sources of blood, 1999 ( n = 35 lots) 3 1  13 NA NA Blood banks, 1998 ( n = 18 lots, 340 680 donors) 17* 12  22 NA NA Individual Samples American Red Cross blood banks, 2000 ( n = 645) 5.6 1.9  52.3 4.6 4.3  4.8 Elderly ( 65  96 years), 2000 ( n = 238) NA 1.4  16.7 4.2 3.9  4.5 Children ( 2  12 years), 1995 ( n = 598) 5.6 1.9  56.1 4.9 4.7  5.1 * PFOA detected in about 1/ 3 of the pooled samples but quantifiable in only 2.

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

Individual blood samples from 3 different age populations were recently analyzed for PFOA and other fluorochemicals using high pressure liquid chromatography/ electrospray tandem mass spectrometry ( HPLC/ ESMSMS) ( Olsen, et al., 2002a, 2002b, 2002c). The studies' participants included adult blood donors, an elderly population participating in a prospective study in Seattle, WA, and children from 23 states participating in a clinical trial. Overall, the PFOA geometric means were similar across all 3 populations ( 4.6 ppb, 4.2 ppb, and 4.9 ppb, respectively). The geometric means and 95% tolerance limits ( the exposure below which 95% of the population is expected to be found) and their upper bounds were comparable across all 3 studies. However, the upper ranges for the children and adults were much higher than for the elderly population. It is not clear whether this is the result of geographic differences in PFOA levels or some other factor. It should be noted that PFOS and PFOA were highly correlated in all three studies ( r = .63, r = .70, and r = .75) and that PFOA did not meet the criteria for a log normal distribution 44 based on the Shapiro Wilk test in any of the studies. However, the data appeared to have a log normal distribution and therefore geometric means were calculated. The authors suggest that it may be due to the greater proportion of subjects with values less than the lower limit of quantitation ( LLOQ); however, only 12 of the 1481 total samples were below the LLOQ. In those instances where a sample was measured below the LLOQ, the midpoint between zero and the LLOQ was used for calculation of the geometric mean. The details of each study are provided below.

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

The geometric mean serum PFOA level was 4.6 ppb. The range was < lower limit of quantitation ( 1.9 ppb) to 52.3 ppb. Only 2 samples were less than the LLOQ. Males had significantly higher ( p < .05) geometric mean PFOA levels than females ( 4.9 ppb vs. 4.2 ppb). Age was not an important predictor of adult serum fluorochemical concentrations. When stratified by geographic location, the highest geometric mean for PFOA was in the samples from Charlotte, NC ( 6.3 ppb, range: 2.1   29.0) and the lowest from Portland ( 3.6 ppb, range: 2.1   16.7). The highest individual value was reported in Hagerstown ( 52.3 ppb).

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

The geometric mean of PFOA for all samples was 4.2 ppb ( 95% CI, 3.9   4.5). The range was 1.4   16.7 ppb. Only 5 samples were less than the LLOQ of 1.4 ppb. There was no significant ( p < .05) difference in geometric means for males and females. In simple linear regression analyses, age was negatively ( p < .05) associated with PFOA in elderly men and women. In bootstrap analyses, the mean of the 95% tolerance limit for PFOA was 9.7 ppb with an upper 95% confidence limit of 11.3 ppb. PFOS and PFOA were highly correlated ( r = .75) in this study.

A sample of 598 children, ages 2 12 years old, participating in a study of group A streptococcal infections, was analyzed for serum PFOA levels ( Olsen, et al., 2002c). The samples were collected in 1994 1995 from children residing in 23 states and the District of Columbia. PFOA did not meet the criteria for a log normal distribution based on the Shapiro Wilk test. The authors suggest that it may be due to the greater proportion of subjects with values < LLOQ for PFOA; however, only 5 samples were less than the LLOQ of 1.9 ppb. The geometric mean of PFOA for all of the participants was 4.9 ppb ( 95% CI, 4.7   5.1). The range was 1.9 to 56.1 ppb. Male children had significantly ( p<. 01) higher geometric mean serum PFOA levels than females: 5.2 ppb and 4.7 ppb, respectively. In simple linear regression analyses, age was significantly ( p < 45 .05) negatively associated with PFOA in both males and females. When stratified by age, the geometric mean of PFOA was highest at age 4 ( 5.7 ppb) and lowest at age 12 ( 3.5 ppb). Although the data were not reported, a graphical presentation of log PFOA levels for each state by gender looked similar across the states; however, it is difficult to interpret these data without the data and given the limited sample size for each gender/ location subgroup. In bootstrap analyses, the mean of the 95% tolerance limit for PFOA was 10.1 ppb with an upper 95% confidence limit of 11.0 ppb. PFOS and PFOA were highly correlated ( r = .70) in this study. PFOA and PFHS ( perfluorohexanesulfonate) were also correlated ( r = .48).

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

In another study, the PFOA concentration was analyzed in human sera and liver samples ( Olsen et al., 2001d). Thirty one donor samples were obtained from 16 males and 15 females over an 18 month period from the International Institute for the Advancement of Medicine ( IIAM). The average age of the male donors was 50 years ( SD 15.6, range 5 69) and the average age of the female donors was 45 years ( SD 18.5, range 13 74). The causes of death were intracranial hemorrhage ( n = 16 or 52%), motor vehicle accident ( n = 7 or 23%), head trauma ( n = 4 or 13%), brain tumor ( n = 2 or 6%), drug overdose ( n = 1 or 3%) and respiratory arrest ( n = 1 or 3%). Both serum and liver tissue were obtained from 23 donors; 7 donors contributed liver tissue only and 1 donor contributed serum only. Serum samples were obtained from 5 ml of blood; liver samples consisted of 10 g of tissue. Samples were frozen at IIAM and shipped frozen to 3M for analysis. Samples were extracted using an ion pairing extraction procedure and were quantitatively assayed using HPLC ESMSMS and evaluated versus an unextracted curve. Extensive matrix spike studies were performed to evaluate the precision and accuracy of the extraction procedure. Serum values for PFOA ranged from < LOQ (< 3.0)   7.0 ng/ mL. Assuming the midpoint value between zero and LOQ serum value for samples < LOQ, the mean serum PFOA level was 3.1 ng/ mL with a geometric mean of 2.5 ng/ mL. No liver to serum ratios were provided because more than 90% of the individual liver samples were < LOQ.

5.0 Preliminary Risk Assessment For this preliminary risk assessment, a margin of exposure ( MOE) approach was used to describe the potential for developmental toxicity associated with exposure to PFOA and its salts. The MOE is calculated as the ratio of the NOAEL, LOAEL, or BMDL for a specific endpoint to the estimated human exposure level. The MOE does not provide an estimate of population risk, but simply describes the relative " distance" between the exposure level and the NOAEL, LOAEL, or BMDL.

For many risk assessments, the MOE is calculated as the ratio of the administered dose from the animal toxicology study to the estimated human exposure level. The human exposure is estimated from a variety of potential exposure scenarios, each of which requires a variety of 46 assumptions. A more accurate estimate of the MOE can be derived if measures of internal dose are available for humans and the animal model. In this preliminary risk assessment, serum levels of PFOA, which is a measure of internal dose, were available for the animal toxicology studies and from human biomonitoring studies. Thus, internal dose was used for the calculation of MOEs in this assessment.

5.1 Selection of Developmental Endpoints As stated in the section entitled " Scope of the Assessment", the purpose of this preliminary assessment was to determine the potential of developmental toxicity associated with exposure to PFOA and its salts. It was therefore necessary to determine which endpoints from the animal toxicology studies would be relevant for this assessment. The selection of developmental endpoints for this assessment was based on the Agency's Developmental Toxicity Risk Assessment Guidelines ( EPA, 1991). According to the guidelines, the period of exposure for developmental toxicity is prior to conception to either parent, through prenatal development and continuing until sexual maturation. In contrast, the period during which a developmental effect may be manifested includes the entire lifespan of the organism.

Several oral prenatal developmental toxicity studies of APFO have been conducted in rats and rabbits. A summary of the exposure duration and the LOAELs and NOAELs are presented in Table 4.

Table 4. Summary of Oral Prenatal Developmental Toxicity Studies Species Exposure Duration Time Endpoints Assessed LOAEL ( mg/ kg/ day) NOAEL ( mg/ kg/ day) Reference Sprague Dawley rat ( n= 22/ group) GD 6 15 GD 20 none 150 Gortner, 1981 Sprague Dawley rat ( n= 25/ group) GD 6 15 GD 20 none 100 Staples, 1984 Sprague Dawley rat ( n= 12/ group) GD 6 15 PND 35 none 100 Staples, 1984 New Zealand white rabbit ( n= 18/ group) GD 6 18 GD 29 50* 5 Gortner, 1982 *  there was a dose related increase in a skeletal variation, extra ribs or 13th rib, with statistical significance at the high dose group ( 38% at 50 mg/ kg/ day, 30% at 5 mg/ kg/ day, 20% at 1.5 mg/ kg/ day, and 16 % at 0 mg/ kg/ day). 47 In addition, developmental effects were observed in the oral two generation reproductive toxicity study that was conducted in Sprague Dawley rats ( York, 2002). For selection of the developmental endpoints from this study, attention was focused on effects that were noted during the period of developmental exposure. Thus, only effects that occurred up to sexual maturation were considered relevant for this preliminary risk characterization. In the high dose group administered 30 mg/ kg/ day APFO, there was a reduction in F1 mean body weight on a litter basis during lactation ( sexes combined). For F1 females, there was a significant increase in mortality mainly during the first few days after weaning, and a significant delay in the timing of sexual maturation in the 30 mg/ kg/ day group. For F1 females, the LOAEL for developmental toxicity was considered to be 30 mg/ kg/ day, and the NOAEL was 10 mg/ kg/ day. For F1 males in the 30 mg/ kg/ day group, there was an increase in mortality mainly during the first few days after weaning, a significant delay in the timing of sexual maturation, and a significant reduction in mean postweaning body weight that began on day 8 and continued through the duration of the study. At 10 mg/ kg/ day, mean body weights were significantly reduced beginning on day 36 postweaning and continuing through the duration of the study. For F1 males, the LOAEL for developmental toxicity was considered to be 10 mg// kg/ day, and the NOAEL was 3 mg/ kg/ day. The NOAEL for the F2 generation offspring was considered to be 30 mg/ kg/ day. No treatmentrelated effects were observed at any doses tested in the study. However, it should be noted that the F2 pups were sacrificed at weaning, and thus it was not possible to ascertain the potential post weaning effects that were noted in the F1 generation.

The database that is available to examine the potential developmental toxicity associated with oral exposure to APFO thus consists of two prenatal studies in Sprague Dawley rats, a prenatal study in New Zealand white rabbits, and a two generation reproductive toxicity study in Sprague Dawley rats. Since no developmental toxicity was noted in the prenatal studies in Sprague Dawley rats these studies were not considered for the calculation of the MOEs. The effects noted in the prenatal study in New Zealand white rabbits and in the two generation reproductive toxicity study in Sprague Dawley rats were considered important for the calculation of the MOEs.

5.2 Use of Serum Levels as a Measure of Internal Dose for Humans Serum levels of PFOA were available from human biomonitoring studies. These provide a measure of total human exposure and a measure of internal dose. The populations that were considered relevant for assessing the potential for developmental toxicity included children and women of child bearing age. Estimates of general human population exposure were available from recent analyses of individual serum samples from a group of children ( 2 12 years) and adults ( 20 69 years). Individual serum data were also available from a recent analysis of a group of elderly adults ( 65 96 years). However, these data were not included in this analysis given that the concerns are for women of child bearing age. The data obtained from the pooled blood samples from blood banks and commercial sources were not used in the calculation of the MOEs given the limitations of these data.

A summary of the human serum levels of PFOA that were considered in the calculation of MOEs is provided in Table 5. The arithmetic means and ranges are presented in order to display 48 the higher end of the range of PFOA serum levels in a small segment of these populations. The geometric means and 95% confidence intervals are presented because the serum levels represent what appears to be a log normal distribution. Gender specific data were available for the geometric mean and range, but not for the arithmetic mean. Since the geometric means for the males and females were very similar ( 4.2 ppb for females and 4.9 ppb for males), the value used in the MOE calculation was the mean for sexes combined which was 4.6 ppb. The value for sexes combined was used so that MOEs could be calculated using both the arithmetic and geometric means. The use of the value of 4.6 ppb versus the gender specific levels of 4.2 and 4.9 ppb has minimal impact on the resulting MOE in this preliminary risk assessment.

Table 5. Summary of Levels of PFOA in the Serum of Human Populations Population Arithmetic Mean Range Geometric Mean 95% CI Adults ( 20  69 years, American Red Cross blood banks, 2000, n= 645) 5.6 ppb 1.9  52.3 ppb 4.6 ppb 4.3  4.8 ppb Children ( 2 12 years, 1995, n= 598) 5.6 ppb 1.9  56.1 ppb 4.9 ppb 4.7  5.1 ppb 5.3 Use of Serum Levels as a Measure of Internal Dose for Animal Studies Serum levels of PFOA were available as a measure of the internal dose of humans. Thus, only those animal studies with serum levels of PFOA were considered for the calculation of the MOEs. Since no information on serum levels was available for the prenatal developmental toxicity study in New Zealand white rabbits, MOEs were not calculated for the endpoints from this study. Serum levels were available for the two generation reproductive study in Sprague Dawley rats. Therefore, this study was used for the calculation of the MOEs.

In the two generation reproductive toxicity study in Sprague Dawley rats, the LOAEL for developmental toxicity was considered to be 30 mg/ kg/ day, and the NOAEL was 10 mg/ kg/ day for the F1 females. For the F1 males, the LOAEL for developmental toxicity was considered to be 10 mg/ kg/ day, and the NOAEL was 3 mg/ kg/ day. Serum levels of PFOA were not measured in the F1 animals. Serum levels were available for F0 animals in the control, 10 and 30 mg/ kg/ day groups. Serum levels were not measured in the 1 and 3 mg/ kg/ day groups. The serum levels were measured in the F0 males at the end of the cohabitation period, while serum levels were measured on lactation day 22 in the F0 females. For both sexes, the serum levels were measured 24 hours after the administration of the last dose.

For this preliminary risk assessment, the serum levels for the F0 animals were used to provide an estimated range of potential serum levels in the F1 animals. The following approach was employed. It was reasoned that if prenatal and/ or lactational exposures were important in 49 eliciting the developmental effects, then the serum levels in the dam ( i. e. F0 females who were being administered APFO) would be the most representative of the serum levels in the F1 pups. The serum levels in the F0 males would not be informative. It was further reasoned that if postweaning exposures were important then the serum levels for the F0 males would be the most appropriate estimate for the F1 males, and similarly the serum levels in the F0 females would be the most appropriate estimate for the F1 females.

If prenatal and/ or lactational exposures were important, several other factors had to be considered. First, serum levels were measured in the F0 females 24 hours after dosing. Given that the serum half life in female rats is less than 24 hours, this value would represent the low end of exposure. Since the peak exposure of the F0 females is not known, it was reasoned that it was unlikely that the peak exposure to the F0 females was higher than the serum level in the F0 males in the same dose group since the serum half life of PFOA in male rats is 4.4  9 days, and therefore with a daily dosing regime they would tend to accumulate PFOA. Therefore, the serum levels in the F0 males and females in the 10 mg/ kg/ day were used to provide a range of values for the calculation of the MOEs; the serum levels were 51.1 and 0.37 ppm in the F0 males and females, respectively. The serum levels of the F0 females probably represent a low internal dose for the F1 animals and the serum levels of the F0 males probably represent a high internal dose for the F1 animals.

For the scenario where postweaning exposures are important, the serum levels in F0 females would be the most appropriate estimate of the serum levels in F1 females. For F1 females, the same rationale was applied for this scenario as was applied to the previous scenario for prenatal and/ or lactational exposures. Therefore, the serum levels in the F0 males and females in the 10 mg/ kg/ day were used to provide a range of values for the calculation of the MOEs.

Similarly, for the scenario where postweaning exposures are important, the serum levels in the F0 males would be the most appropriate estimate of the serum levels in F1 males. For this assessment, the LOAEL for developmental effects in the F1 males was considered to be10 mg/ kg/ day and the NOAEL was considered to be 3 mg/ kg/ day. However, serum levels for the F0 males were only measured in the 10 and 30 mg/ kg/ day groups; there was no information available for the 3 mg/ kg/ day group. In addition, it was not possible to extrapolate the serum levels to the lower administered doses as the values appear to have reached a plateau at 10 and 30 mg/ kg/ day ( 51.1 and 45.3 ppm, respectively) and are not linear. Therefore, for this preliminary assessment the serum levels from the F0 males in the 10 mg/ kg/ day group were used in the calculation of the MOEs.

5.4 Calculation of MOEs The human populations of concern for this preliminary assessment are women of child bearing age and children. As stated in section 5.2, the serum data from the American Red Cross study included both men and women, ages 20 69. As explained in section 5.2, since the data were not consistently reported separately for each gender and since the geometric means were very similar for males and females, the means for the sexes combined are used as a surrogate for women of child bearing age. The MOEs were calculated by dividing the serum values for the F0 female 50 and male rats in the 10 mg/ kg/ day group ( 0.37 ppm and 51.1 ppm, respectively) in the two generation reproductive toxicity study by the American Red Cross blood samples and children's samples presented in Table 5. The MOEs for potentially exposed populations are presented in Table 6.

It is important to note that MOEs that were calculated from the serum levels in the F0 female and male rats provide a means to bracket the low and high ends of reported experimental exposures. This is an unusual situation in that MOE estimates, which typically represent point estimates, are described here as a range of potential values due to uncertainties in the rat serum data. This situation arises from the fact that the available data do not allow selection of a particular departure point for the MOE calculations. It is likely that the MOEs calculated using the F0 female rat serum level are lower than what would be anticipated in the human population, and it is likely that MOEs calculated using the F0 male rat serum level are higher than what would be anticipated in the human population. As uncertainty around the rat serum values decreases the end brackets are likely to shift towards the middle of the current range. Therefore, MOE values presented in Table 6 should not be interpreted as representing the range of possible MOEs in the US population. It is likely that when more extensive rat kinetic data are available, the resultant, refined estimated range of MOEs will constitute a narrower subset of the range presented here.

Table 6. MOE Calculations for Potentially Exposed Populations Using F0 Rat Serum Values and Human Serum Values Human Population MOE values calculated using rat serum values from the 2 generation reproductive study1 Women of Childbearing Age2 Arithmetic mean Geometric mean 66  9125 80  11,109 Children3 Arithmetic mean Geometric mean 66  9125 75  10,429 1Estimated MOE values are bracketed by the serum level concentrations in the F0 females ( 0.31 ppm) and the F0 males ( 51.1 ppm). 2The American Red Cross serum samples were used as an estimate of the arithmetic and geometric means in women of child bearing age. 3The samples from the Children's study were used as an estimate of the arithmetic and geometric means. 51 The MOE that was calculated at the low end of the range using the rat female F0 serum levels and the arithmetic mean for the adults is 66, while the MOE that was calculated using the geometric mean for the adults is 80. The MOEs that were calculated at the high end of the range using the serum levels of the rat F0 males and the arithmetic or geometric means for the adults range from approximately 9,000 11,000. If the upper and lower values of the human serum levels are used in the MOE calculation with the rat F0 females, the resulting MOEs are 195 and 7. If the upper and lower values of the human serum levels are used in the MOE calculation with the rat F0 males, the resulting MOEs are 26,895 and 911.

The MOEs for children were calculated by dividing the serum values for the rat F0 females and males in the 10 mg/ kg/ day group ( 0.37 ppm and 51.1 ppm, respectively) by the biomonitoring values presented in Table 5. The MOE that was calculated at the low end of the range using the rat female F0 serum levels and the arithmetic mean for the children is 66, while the MOE that was calculated using the geometric mean for the children is 75. The MOEs that were calculated at the high end of the range using the serum levels of the rat F0 males and the arithmetic or geometric means for the children range from approximately 9,000 10,400. If the upper and lower values of the childrens' serum levels are used in the MOE calculation with the F0 females, the resulting MOEs are 195 and 6.6. If the upper and lower values of the childrens' serum levels are used in the MOE calculation with the F0 males, the MOEs are 26,895 and 911.

5.5 Uncertainties in the Preliminary Risk Characterization Some of the uncertainties encountered in this preliminary risk assessment are common for many risk assessments. One such example pertains to the choice of the animal model. In this preliminary assessment, serum levels were not available for the oral rabbit prenatal developmental toxicity study and therefore this study was not used in the calculation of the MOEs. It is not known whether rabbits are a more appropriate animal model than rats.

Other uncertainties that are common for many assessments have been avoided in this preliminary assessment through the use of a measure of internal dose of PFOA for both humans and the rat animal model. This approach has avoided many of the pitfalls encountered in trying to estimate human exposure levels through the application of various models and assumptions. Unlike many environmental chemicals where it is only hypothesized that humans are exposed, serum data from humans gives direct evidence that exposure to PFOA has occurred in the general public, and provides a measure of internal dose. Although, it is not known when or how exposure has occurred, this is less of an issue for PFOA given its long half life in humans. Therefore, the MOEs calculated in this preliminary risk assessment can be considered to be a more accurate comparison of the relative " distance" between the exposure level and the NOAEL than if administered dose had been used.

Although the use of serum levels as a measure of internal dose introduces uncertainties that are unique to this assessment, many of these have been accounted for through the use of a range of MOEs. For example, one area of uncertainty pertains to the use of serum data for the F0 animals as estimates of serum levels in the F1 animals. As noted in the previous section, it is not known whether the effects on postweaning mortality, body weight, or age of sexual maturation were due 52 to prenatal exposures, lactational exposures, postweaning exposures, or a combination of one or more of these exposure periods. In most risk assessments, no attempt is made to determine which of these exposure periods is important. A major strength of this preliminary assessment is that each of these exposure periods was considered in order to determine the appropriateness and uncertainties associated with the use of the serum levels from the F0 animals.

It was reasoned that if prenatal and/ or lactational exposures were important then the serum levels in the F0 females would be the most appropriate estimate for the F1 animals. If postweaning exposures were important then the serum levels for the F0 males would be the most appropriate estimate for the F1 males, and similarly the serum levels in the F0 females would be the most appropriate estimate for the F1 females. It was not possible to make a " direct" estimate of F1 serum levels from the serum levels in the F0 females for the prenatal and/ or lactational exposure scenario for several reasons. First, as noted in the previous sections, there is a gender difference in the elimination of PFOA in rats. In female rats, estimates of the serum half life range from 1.9 to 24 hours, while in male rats estimates of the serum half life range from 4.4 to 9 days. In female rats elimination of PFOA appears to be biphasic; a fast phase occurs with a half life of approximately 2 4 hours while a slow phase occurs with a half life of approximately 24 hours. In the two generation reproductive toxicity study, the animals were dosed by gavage once daily. The serum levels were measured 24 hours after dosing. Thus, the values obtained for the F0 females represent the low end of exposure. With no knowledge of the peak exposures, it was reasoned that it was unlikely that the peak exposure would be higher than the serum level in the F0 males in the same dose group since they would tend to accumulate PFOA with a daily dosing regime. Therefore, the strategy that was employed in this assessment was to use the MOEs that were calculated from the serum levels in the F0 females and males as a range or as a means to bracket the low and high ends of exposure. Thus, it is likely that the MOEs that were calculated using the F0 female serum levels are probably too low, while the values calculated from the F0 male serum levels are probably too high.

As noted above, the serum levels from the F0 females would be the most appropriate estimate for the F1 females if postweaning exposures were important. Given the issues associated with the gender difference in elimination of PFOA in rats and the timing of serum collection in the F0 females, the same logic which was applied for the prenatal and lactational exposures was also used for this scenario. Again, the MOEs that were calculated from the serum levels in the F0 females and males were viewed as a range or as a means to bracket the low and high ends of potential exposure.

Similarly, the serum levels from the F0 males would be the most appropriate estimate for the F1 males if postweaning exposures were important. As stated above, the LOAEL for developmental effects in the F1 males was considered to be10 mg/ kg/ day and the NOAEL was considered to be 3 mg/ kg/ day. However, serum levels for the F0 males were only measured in the 10 and 30 mg/ kg/ day groups; there was no information available for the 3 mg/ kg/ day group. In addition, it was not possible to extrapolate the serum levels to the lower administered doses as the values appear to have reached a plateau at 10 and 30 mg/ kg/ day ( 51.1 and 45.3 ppm, respectively) and are not linear. Thus, if postweaning exposures were important, the MOEs based on the serum levels in the F0 males would be too high and underestimate potential risks. Again, this 53 uncertainty has been accounted for in this preliminary assessment through the calculation of a range of MOEs.

In the scenarios presented above, the serum levels of the F0 males were viewed as an overestimate given that male rats tend to accumulate PFOA given the long half life and the fact that saturation had apparently occurred. One line of reasoning that would support the statement that the serum levels from the F0 males are probably too high an estimate of peak exposure in the F0 females is provided by consideration of the serum half life of PFOA in female rats. Four studies have examined serum half life in a variety of rat strains; these studies have employed different routes of exposure and different methods for quantitating serum PFOA levels. Two of these studies were most relevant to the rat strain and route employed in the two generation reproductive toxicity study. Ophaug and Singer ( 1980) estimated a serum half life of 10 hours after administering female Holtzman rats a gavage dose of 8 mg/ kg, while Vanden Heuvel et al. ( 1991b) estimated a serum half life of 4 hours after administering female Harlan Sprague Dawley rats an intraperitoneal dose of 4 mg/ kg. Using a serum half life value of 10 hours would yield a peak internal dose of approximately 1.85 ppm, while a serum half life of 4 hours would yield a peak internal dose of approximately 24.7 ppm. While neither of these values may be absolutely correct given that serum half life has not been reported for the rat strain and dosing conditions used in the two generation reproductive toxicity study, both values do support the statement that the serum levels in the F0 males are probably too high ( 1.85 and 24.7 ppm versus 51.1 ppm). Thus the resultant MOEs using the F0 male serum values probably underestimate the potential risks.

It is possible that lactational and postweaning exposures to the F1 pups may have been higher than the exposures to the F0 females due to the time required for maturation of the clearance mechanism. First, the clearance is under hormonal control and these hormones do not reach adult levels until puberty. Second, recent studies have examined the role of three organic anion transporters ( OAT), OAT1, OAT2, and OAT3, in the urinary elimination of PFOA in the rat. Kudo et al. ( 2002) has provided evidence that OAT2 and OAT3 may be involved. A study of developmental and gender specific influences on the expression of rat OATs in the kidney has shown that at birth all OAT mRNA levels are low ( Buist et al., 2002). Renal OAT1 expression approaches adult level at 30 days, where at day 40 and 45 OAT1 levels were greater in males than females. OAT2 expression was minimal through day 30 but increased dramatically only in females at day 35. OAT3 expression matured the earliest and reached adult levels at 10 days. If these OATs are important, then this developmental profile would suggest that the PFOA clearance in juvenile female rats is less than in adult female rats. Therefore, the serum levels in the females prior to sexual maturation may be higher than in the adult, and the resultant MOEs would also be higher. However, it is important to note that the MOEs in this situation would still fall within the range of MOEs that were calculated in this preliminary assessment.

There are several other uncertainties unique to this assessment. The first pertains to the choice of developmental endpoints. This preliminary assessment utilized endpoints from the two generation reproductive toxicity study that could be directly attributed to developmental exposures. Another way to ascertain potential developmental effects would be to compare the magnitude and dose levels of the systemic toxicity that was observed in the F1 animals 54 ( developmental and adult exposures) at the end of the study period with those systemic effects noted in the F0 animals ( adult exposures only). For example, if organ weight changes were greater in F1 than F0 animals or if changes occurred at lower doses in the F1 animals, this may be indicative of the importance of developmental exposures. In the F0 males, there were significant reductions in the absolute weights of the left and right epididymides, left cauda epididymis, seminal vesicles ( with and without fluid), prostate, pituitary, left and right adrenals, spleen, and thymus at 30 mg/ kg/ day. In the F1 males, the effects on many of these organ weights occurred at lower doses. This may indicate that the developmental exposures were important, and that the LOAEL and NOAEL are lower than the values used in this preliminary assessment. If true, this preliminary assessment would have underestimated potential risks.

Another area of uncertainty pertains to the use of the serum levels as a measure of internal dose. PFOA behaves differently from many other persistent environmental contaminants in that it is not stored in adipose tissue. It is not clear whether it binds to proteins or other macromolecules, but it is clear, based on animal studies, that PFOA enters enterohepatic circulation. Animal studies indicate that PFOA mainly partitions to blood serum and the liver. In this preliminary assessment, serum levels were used primarily because this was the only information available for humans. However, it is possible that area under the curve, liver levels, the ratio of serum to liver levels, or some other measure may be more appropriate dose metrics.

In addition, the apparent biphasic elimination of PFOA in female rats raises an important issue, whether the observed effects are due to the kinetics associated with the fast or slow elimination phase in the females. The data currently available do not allow development of this potentially important issue. In this preliminary assessment, the serum values for the F0 males were used to provide an estimate of the peak exposure of the F0 females. The MOEs that were then calculated from the serum levels in the F0 females and males provide a means to bracket the high and low ends of exposure. If the effects are associated with the slow elimination phase, then the serum levels in the F0 females and the resultant MOEs would be more realistic.

Another area of uncertainty pertains to the differences in the serum half life in humans and rats. In humans, the serum half life is years, while in the rat it is less than 24 hours in females and 105 hours in males. To date, there is no evidence of a gender difference in the elimination of PFOA in humans. Thus, humans appear to be more similar to male rats in that they will tend to accumulate PFOA and will have a more continuous internal exposure. It is not known how this would impact potential risks of developmental toxicity in humans.

Finally, there is some uncertainty regarding the use of the human biomonitoring data. Although the available data include a range of populations with various demographics, there may be some populations that may not be represented. Since it is unknown how the human exposures are occurring, proximity to a manufacturing plant may be a factor in exposure. However, populations living near the plants were not sampled. Therefore, it is possible that PFOA serum levels may be underestimated for certain portions of the U. S. population. The children's sample was derived from blood collected in 1994/ 1995; therefore, it may not reflect the current status of PFOA in children's blood. It is not clear how PFOA may affect more sensitive subpopulations or if their exposures would vary. 55 6.0 Overall Conclusions This preliminary risk assessment focused on the potential risks for developmental toxicity associated with exposure to PFOA and its salts. Concerns for developmental toxicity were raised from the results of a rat two generation reproductive toxicity study of APFO . In this study, there was a reduction in F1 mean body weight on a litter basis during lactation ( sexes combined). Postweaning mortality and delayed sexual maturation were noted in F1 females administered 30 mg/ kg/ day APFO; the NOAEL for developmental effects for F1 females was 10 mg/ kg/ day. Postweaning mortality, delayed sexual maturation and a significant reduction in postweaning body weights were noted in F1 males at 30 mg/ kg/ day, and a significant reduction in postweaning body weight was noted at 10 mg/ kg/ day. For F1 males, the LOAEL for developmental effects was 10 mg/ kg/ day and the NOAEL was 3 mg/ kg/ day.

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