+-------------------+--------------------------------------------------+
| >                 | October 29, 2003                                 |
|  ![](media/image1 |                                                  |
| .png){width="1.34 | Office of Environmental Information Docket       |
| 30555555555554in" |                                                  |
| > height="1.588   | Environmental Protection Agency                  |
| 1944444444445in"} |                                                  |
| >                 | Mailcode 28220T                                  |
| > *A*             |                                                  |
| >                 | 1200 Pennsylvania Avenue, NW                     |
| > *Council*       |                                                  |
| >                 | Washington, D.C. 20460                           |
| > *of the*        |                                                  |
| >                 | Re: Comments of the Chlorine Chemistry Council   |
| > *American*      | on EPA's Framework for the Application of the    |
| >                 | Toxicity Equivalence Methodology for             |
| > *Chemistry*     | Polychlorinated Dioxins, Furans, and Biphenyls   |
| >                 | in Ecological Risk Assessment,                   |
| > *Council*       |                                                  |
| >                 | Docket ID No. ORD-2003-0002                      |
| > [1300 Wilson    |                                                  |
| > B               | The Chlorine Chemistry Council (CCC) appreciates |
| lvd.]{.smallcaps} | this opportunity to comment on the U.S.          |
| >                 | Environmental Protection Agency's Framework for  |
| > [Arlington, VA  | the Application of the Toxicity Equivalence      |
| >                 | Methodology for Polychlorinated Dioxins, Furans, |
|  USA]{.smallcaps} | and Biphenyls in Ecological Risk Assessment      |
| >                 | (Framework). CCC, a business council of the      |
| > [2              | American Chemistry Council, is dedicated to      |
| 2209]{.smallcaps} | addressing public policy issues affecting the    |
| >                 | chlorine industry. Because EPA's Framework has   |
| > [▲]{.smallcaps} | the potential to affect the chlorine industry by |
| >                 | changing the method by which risk assessments    |
| > [Tel:           | for dioxins and dioxin-related compounds are     |
| > 703-741-        | conducted, CCC has a strong interest in the      |
| 5850]{.smallcaps} | Framework.                                       |
| >                 |                                                  |
| > [Fax:           | The U.S. Environmental Protection Agency (EPA)   |
| > 703-741-        | has proposed to refine the current method by     |
| 6850]{.smallcaps} | which it calculates toxicity for purposes of its |
| >                 | ecological risk assessments of dioxins and       |
| > [703-741-       | dioxin-like compounds. CCC believes there are    |
| 6084]{.smallcaps} | both valuable guidance points as well as a       |
| >                 | number of problems inherent with the proposed    |
| > www.c3.org      | Framework. These comments outline a number of    |
|                   | those problems and address specific issues as    |
|                   | they pertain to the Framework and the            |
|                   | application of TEFs to dioxins and dioxin-like   |
|                   | compounds. In addition, CCC supports the         |
|                   | comments of the American Chemistry               |
|                   | CouncilPolychlorinated Biphenyls Panel (ACC PCB  |
|                   | Panel).                                          |
|                   |                                                  |
|                   | ## I. Execu                                      |
|                   | tive Summary {#i.-executive-summary .unnumbered} |
|                   |                                                  |
|                   | EPA should not adopt the currently proposed      |
|                   | Framework for application in assessing           |
|                   | ecological risks to exposures of dioxins and     |
|                   | dioxin-like compounds. As described more fully   |
|                   | below, inherent limitations exist in the         |
|                   | toxicity equivalence factor (TEF) methodology    |
|                   | that do not necessarily improve ecological risk  |
|                   | assessment (ERA) methods already in place.       |
|                   | Indeed, contrary to what is stated in the        |
|                   | Framework, the use of TEFs does not necessarily  |
|                   | reduce uncertainty. For example, TEF estimates,  |
|                   | in most instances, are not based on population   |
|                   | effects that are the hallmark measures in ERA.   |
|                   | TEFs are usually rounded-up a half-order of      |
|                   | magnitude, thus introducing uncertainty into     |
|                   | ecological risk assessments that rely on TEFs.   |
|                   | This uncertainty is evidenced further by the     |
|                   | fact that many TEFs are based solely on a study  |
|                   | of a single compound rather than a complex       |
|                   | environmental mixture of dioxins.                |
|                   |                                                  |
|                   | #                                                |
+-------------------+--------------------------------------------------+

The Framework should be limited to a screening-level risk assessment
tool. At a minimum, the Framework should quantify the uncertainty for
each dioxin and furan TEF so that it is apparent in any ecological risk
assessment that relies upon the TEF methodology. For example, full
disclosure of the strengths and weaknesses for the current TEF estimates
should be provided in the Framework so that the reader can be informed
on the limited data used to establish TEFs for dioxins and certain
dioxin-like congeners.

TEF methodology assumes that the combined effects of dioxin-like
congeners are additive. Yet, the only mechanistic similarity EPA relies
on to support an additive approach is Ah receptor binding. Additionally,
additivity in PCDD and PCB mixtures has not been demonstrated across
congeners and endpoints in animal studies (Harper et al., 1995; Safe,
1990; Starr, 1997). In fact, little has been done to elucidate the
mechanism of action, or even the mode of action, for dioxins and
dioxin-like substances. Therefore, it is unwarranted to assume that the
toxicity of such mixtures can be predicted by summing the TEFs for the
individual congeners.

The Framework does allow for flexibility in using TEFs, or related data,
in the event superior scientific data exist for conducting ERAs, e.g.,
the common sense perspective stated in the Framework. This Framework
document contains a number of useful perspectives on using TEFs versus
relative potency factors (RPFs) and checklist criteria for properly
conducting an ERA. However, CCC recommends that the Framework discuss
more prominently the need to conduct actual population and body burden
studies in order to experimentally validate modeled body burdens and
hazard indices that rely on TEF methodology.

Finally, the Framework should also include summary information in the
Preface to provide quicker access to the salient aspects contained in
the Framework. For example, two recent summary papers given at the 2003
Dioxin Conference in Boston concisely addressed the critical elements of
this Framework, as well as a decision process for selecting the
appropriate TEF, RPF, or other toxicity data that most appropriately
meet the need of site-specific ERA questions (Cook and Henry, 2003;
Henry et al., 2003).

**II. Conceptual and Technical Criticisms of the Framework's Reliance on
the TEF Methodology**

> > A. [The TEF Methodology Provides No Marked Improvement Over Current
> > Ecological Risk Assessment Technologies.]{.underline}

The Framework fails to provide sufficient evidence that the TEF
methodology will significantly improve how ecological risk assessments
are currently conducted, improve the past record of remediation
activities, or impact the recovery of impacted ecosystems. The Framework
does not cite any consensus conference proceedings, or academic,
industry, and government scientists, concluding that the TEF methodology
will provide an improvement over how current ecological risk assessments
are performed.

TEFs likely differ for different endpoints, doses, species, and types of
mixtures. Yet the Framework relies on many assumptions, namely that each
compound assigned a TEF elicits the same toxic endpoint (both cancer and
non-cancer), that dose-response curves are parallel, and that the
toxicities of these compounds are additive, regardless of the mixture.

> > B. [The Framework Should be Limited to a Screening-Level Risk
> > Assessment Tool .]{.underline}

The Framework should be characterized as a screening-level risk
assessment tool. Furthermore, the Framework should place greater
emphasis on site-specific studies designed to determine the presence of
adverse population effects. Actual population studies using measurements
of tissue dioxin and dioxin-like activity are more valid.

TEF methodology has not been validated for estimation of
population-level field effects, which is a widely accepted technique for
conducting ecological risk assessments. Nonetheless, the Framework
incorrectly claims: "Use of the toxicity equivalence methodology results
in more precise characterization of AhR mediated stressors and their
potential effects in ecological receptors." Framework at p. 14. This
statement is not supported by appropriate references and thus goes
beyond the scientific evidence linking TEF tissue concentrations modeled
with biomagnification factors and with adverse population effects.
Further, the Framework addresses only endpoints for individuals rather
than populations. This approach is in direct contrast to the strategy
set forth in EPA's guidelines for ecological risk assessment. EPA should
stress that the TEF methodology, as outlined in the Framework, is a
screening tool that assesses potential risks to individuals, rather than
populations.

C. [The TEF Methodology is based on the incorrect assumption that doses
are additive for mixtures containing dioxins, furans, or co-planar PCB
congeners.]{.underline}

TEF methodology is based on the assumption that the doses of dioxins,
dioxin-like congeners and PCBs are additive, i.e., that the toxicity of
the mixture can be derived by adding the doses (modified through the
application of TEFs) of TCDD and dioxin-like substances. Dose additivity
can only occur when substances have the same mechanism of action. Using
a less stringent standard for dose additivity, EPA assumes that
substances share a common mechanism of action or a "common mechanism of
toxicity" when they "share major steps leading to an adverse health
effect following interaction of a substance with biological targets"
(U.S. EPA, 1999; 2000).

A chemical that binds to the Ah receptor and causes any effect is termed
an agonist. Conversely, a chemical that binds but has no effect or
otherwise inhibits the occurrence of an effect is called an antagonist.
Chemicals that bind to a receptor with no adverse
effectantagonistscompete with agonists for sites on receptors. Thus,
while an antagonist occupies the receptor site, an agonist does not
occupy it and therefore has no effect. Although the Ah receptor is
capable of binding with a variety of molecules, the configuration of a
chemical molecule determines whether binding occurs, as well as the
strength of that bind. Should a chemical bind weakly to the Ah receptor
it may be displaced by a competing chemical capable of creating a
stronger bond with that receptor. Therefore, Ah receptor binding may be
as competitive as it is additive.

Furthermore, the Framework states "empirical data support the use of the
additivity concept." Framework at p. 10. This empirical evidence is
insufficient particularly when there are significant financial impacts
of the TEF-derived risk estimates on remediation activities and costs.
CCC disagrees with the statement, "A substantial effort has been made to
test the assumptions of additivity and the ability of the toxicity
equivalence methodology to predict the effects of mixtures of
dioxin-like chemicals," especially when it comes to predicting
population effects. Framework at p. 66. Instead, the Framework should
recommend further experimental studies of mixture effects, preferably on
populations, to determine if the TEF methodology is appropriate and
valid.

> D. [The TEF Methodology and Societal Costs.]{.underline}

TEFs lack the precision necessary for making important and sound
economic decisions with respect to remediation. The rounding effect for
deriving the final TEF estimate can lead to substantial, and possibly
unwarranted, remediation costs. EPA should quantify the uncertainty for
TEFs so that the uncertainty is apparent in any ecological risk
assessment that relies upon TEF methodology. Agency for Toxic Substances
and Disease Registry (ATSDR) recognized the implications of applying
multiple uncertainties to a dioxin risk assessment when it stated, "In
practice, there is a big cost difference in trying to clean up a
hazardous waste site so that the final residue of dioxins is at the 10
ppb, 1 ppb, or 0.1 ppb level. Risk assessment is not a precise science,
and different clean-up levels may be driven by or considered by the
public as artifacts of the application of uncertainty factors. Because
of the limited budget for environmental clean-up, overprotection at one
site may result in lack of funds for another site where the resources
are needed. For every environmental pollutant, health risks, clean-up
benefits, and economical feasibility must be carefully evaluated" (Pohl
et al., 2002). Basing an ecological risk assessment on TEF values
fraught with multiple uncertainties and unproven assumptions can greatly
alter the standards to which a site must be cleaned.

> E. [Limited Studies and Available Sensitivity Data.]{.underline}

Limited studies exist for TEF derivations for fish and birds. A vast
majority of the evidence supporting TEF application in fish has been
done using Salmonids and injection of eggs as a method of chemical
dosing. Despite this limited scientific basis, the Framework clearly
intends for the TEF values to be applied to a wide variety of species.

The available data demonstrate a range of sensitivities that should be
evaluated according to species's sensitivity distributions. This need is
clearly articulated in the following statement: "The relative
sensitivity to dioxin-like toxicity among species that possess the Ah
receptor varies greatly, even within taxonomic class." Framework at p.
18.

Finally, CCC is also concerned about the apparent contradictions in the
Framework. For example, the Framework claims that the TEF approach will
strengthen ERA while reducing uncertainty, yet on page 20 the Framework
admits: "Given the known differences in sensitivity among species and
endpoints, risk assessors should consider the uncertainty introduced
when extrapolating from a species or endpoint for which sensitivity has
been established to a specie or endpoint of unknown sensitivity." Hence,
this Framework should clearly articulate the relative newness of the TEF
approach and its need for substantial improvement that only further
experimental studies can provide.

##  {#section-1 .unnumbered}

## III. Other Technical Issues {#iii.-other-technical-issues .unnumbered}

##### A. Lack of Validation and Lack of Precision for BSAF {#a.-lack-of-validation-and-lack-of-precision-for-bsaf .unnumbered}

The use of biota-sediment accumulation factor (BSAF) values has been
proposed as an empirical relationship to predict the lipid-normalized
residues of dioxins and dioxin-like compounds from organic
carbon-normalized sediment residue of the same compounds:

BSAF = C~t~/C~s~

where, BSAF = biota-sediment accumulation factor

> C~t~ = chemical concentration in the animal's tissues, normalized for
> lipid content (ng-g^-1^/% lipid)
>
> C~s~ = chemical concentration in sediment, normalized for sediment
> organic carbon content (ng-g^-1^/% total organic carbon)

The BSAF approach has been proposed in this Framework as a tool for
estimating toxicity equivalence concentrations (TECs) in fish, bird
eggs, and dietary items from dioxins and dioxin-like compound
concentrations in surface sediment. The Framework uses tables of BSAF
values to estimate TEC values in a variety of organisms, including trout
eggs, gull eggs, and forage fish. Although the Framework correctly
identifies BSAF values as varying with site-specific information such as
food chain position, it neglects other important confounding issues such
as habitat, location, and species (Muir et al., 1992; Lake et al.,
1990); sediment particles, carbon type and amount (Lake et al., 1990;
Boese et al., 1995; Froese et al., 1998); and contaminant concentration
(Ferraro et al., 1990).

EPA researchers (Ferraro et al., 1990) have observed that the BSAF model
provides reasonable estimates only for lipophilic compounds in
organically enriched, highly polluted sediments. The tool, however, was
far less effective in sediments with low organic carbon content and/or
low chemical residue levels. The duration of the contaminant's presence
in the sediment and exposure of biota to the sediment can also influence
the BSAF value, particularly for highly lipophilic materials that may
require a lengthy period to achieve steady state with both sediment and
biological matrices (Hawker and Conner, 1985). In use of a BSAF for a
given material, significant variations in organism residue values (C~t~)
can be observed if a limited number of samples of TOC-normalized
sediment concentrations (C~s~) are used to estimate this contributor of
the BSAF value itself. Variations of as much as 35-fold may result when
single input BSAF values are applied (Froese et al., 1998)

The absence of any references for the listed BSAF values does not allow
for evaluation of the suitability of the values. A comparison of
calculated organism residues, based on BSAF values, to field residues
should be included in the Framework. For instance, there is no
discussion of how the predicted organism TECs compare to field measured
TEC values.

In summary, the use of BSAF values in the Framework incorrectly assumes
that these values do not vary with location, time, or concentration and
that an overall average BSAF value can be calculated for a food web
position. Based on the long list of confounding variables, the data
suggest that BSAF values may be useful for screening-level risk
assessments, but not for higher tier assessments, as significant errors
in estimated chemical residues in biota may result. The range or
uncertainty of BSAF values should be considered in addition to the
average value in interpreting risk calculations, as this will place any
calculated organism residues in a proper context.

> B. [Differences in Species Responsiveness]{.underline}

Elonen et al. (1998) have reported 32- to 100-day chronic fish early
life-stage no-observed effect concentration (NOEC) values for the
following species exposed to 2,3,7,8-TCDD:

> Table 1. Fish species differences in chronic toxicity response upon
> exposure to 2,3,7,8-TCDD (fish egg concentrations, pg/g).
>
> [Species NOEC LOEC LC~10~/LC~50~]{.underline}
>
> Lake herring 175 270^a^ 509/902
>
> Fathead minnow 235 435^a^ 293/539
>
> Channel catfish 385 855^a^ 429/644
>
> Zebrafish 424 2000^a^ 1610/2610
>
> Japanese medaka 455 949^a^ 656/1110
>
> White sucker 848 1220^b^ 1590/1890
>
> Northern pike 1190 1800a 1530/2460
>
> ^a^ Significant decrease in survival compared to controls (p \< 0.05)
>
> ^b^ Significant decrease in growth compared to controls (p \< 0.05)

In addition to the data presented above in Table 1, other fish toxicity
NOEC values for TCDD in fish eggs range from 35 pg/g for lake trout to
125 pg/g for brook trout (Walker and Peterson, 1994). In brief, the
range of fish species differences in functional responses to TCDD is
over a factor of 30 (35 to 1190 pg/g, ww). These fish species
differences are likely due to several factors, including toxicokinetics,
receptor distribution and affinity, and agonistic action on receptor
upon binding (Van den Berg et al., 2000). The issue of toxicokinetics is
an important one, as the congeners vary considerably in their
environmental fate and behavior in biota. The use of tissue
concentrations to establish TEF values can be an important source of
error, in that the chemicals may vary considerably in their tissue
distribution. These differences in tissue distribution may help account
for some of the variability noted between various fish species. Such
varying toxicokinetic factors have not been accounted for in the
assignment of current TEF values for fish. The current TEF values have
generally been developed using data on Salmonidae fish, which tend to be
more sensitive to TCDD toxicity than most other aquatic vertebrates. The
use of TEF values developed using only the sensitive Salmonidae species
may not accurately represent the level of toxic response for other
species because specie differences in toxic response vary by a factor of
more than 30.

In summary, fish species other than Salmonidae have been examined for
dioxins and dioxin-like compound toxicity, including zebrafish, fathead
minnow, bullhead and channel catfish, mosquitofish, guppy, bluegill,
largemouth bass, and yellow perch. Collectively, these studies
demonstrate the general susceptibility of fish to dioxins, but
differences in exposure protocols (flow-through, static, static renewal,
egg injection, and dietary), life stages tested (adult, juvenile,
larval, embryo, and egg), the toxicological endpoints examined
(pathology, growth, mortality, enzyme induction, and development) and
the general lack of tissue-specific congener data, have indicated a lack
of an integrated understanding of various aquatic species and life stage
sensitivity. The current TEF values reflect the limited dataset upon
which development of these values is based.

### IV. Conclusion

Based on the inherent limitations in the TEF methodology, CCC recommends
that EPA not adopt this methodology for application in assessing
ecological risks from exposures to dioxins and dioxin-like compounds.
Instead, [the Framework should be limited to a screening-level risk
assessment tool.]{.underline} At a minimum, the Framework should
quantify the uncertainty for each dioxin and furan TEF, making it
apparent in any ecological risk assessment that relies upon TEF
methodology. EPA should encourage experimental studies to validate the
benefits of the TEF modeling approach and point to the uncertainties in
the TEF methodology, such as the lack of precision for BSAF, differences
in species responsiveness and sensitivity, and mechanistic steps in
receptor binding. Finally, CCC recommends that EPA emphasize the need to
conduct actual population and body burden studies to experimentally
validate modeled body burden and hazard indices based on TEF methodology
used in ecological risk assessments.

Please direct any questions or comments you may have concerning this
submission to Todd Abel (703-741-5856).

Sincerely,

![](media/image2.png){width="1.5in" height="0.5409722222222222in"}

Clifford T. "Kip" Howlett, Jr.

Executive Director,

American Chemistry Council,

Vice President

**VI. References**

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1995. PCB congeners and hexachlorobenzene biota sediment accumulation
factors for Macoma nasuta exposed to sediments with different total
organic carbon contents. *Environ. Toxicol. Chem.* 14: 303-310.

Cook, P.M. and T.R. Henry. 2003. A conceptual model for evaluating
relative potency data for use in ecological risk assessments. Presented
at *Dioxin 2003*, Boston, MA. August.

Elonen et al. 1998. *Environ. Toxicol. Chem.* 17: 472-483.

Ferraro, S.P., H. Lee, R.J. Ozretich, and D.T. Specht. 1990. Predicting
bioaccumulation potential: A test of a fugacity-based model. *Arch.
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Froese, K.L., D.A. Verbrugge, G.A. Ankley, G.J. Niemi, C.P. Larsen, and
J.P. Giesy. 1998. Bioaccumulation of polychlorinated biphenyls from
sediment to aquatic insects and tree swallow eggs and nestlings in
Saginaw Bay, Michigan, USA. *Environ. Toxicol. Chem.* 17: 484-492.

Harper, N., K. Connor, M. Steinberg, and S. Safe. 1995.
Immunosuppressive activity of polychlorinated biphenyl mixtures and
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Hawker, D.W. and D.W. Conner. 1985. Relationships between partition
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Henry, T.R., P.M. Cook, P. Cirone, M. De Vito, B. Duncan, R. Pepin, S.
Schwenk, and S. Wharton. 2003. Framework for application of the toxicity
equivalence methodology for polychlorinated dioxins, furans, and
biphenyls in ecological risk assessment. Presented at *Dioxin 2003*,
Boston, MA. August.

Lake, JL, N.I. Rubinstein, H. Lee, C.A. Lake, J. Heltshe, and S.
Pavignano. 1990. Equilibrium partitioning and bioaccumulation of
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Muir, D.C.G., M.L. Fairchild, and D.M. Whittle. 1992. Predicting
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Pohl, H.R., H.E. Hicks, D.E. Jones, H. Hansen, and C.T. De Rosa. 2002.
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and R.B. Connolly. 1997. Concerns with the use of a toxicity equivalence
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*Organohal. Comp.* 34:91--94.

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64:5796-5799.

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817-820.