# ![](media/image1.jpeg){width="0.5868055555555556in" height="0.9736111111111111in"}

November 6, 2003

(Via E-Mail)

Office of Environmental Information Docket

Environmental Protection Agency (28220T)

1200 Pennsylvania Ave., N.W.

Washington, D.C. 20460

> Re: AF&PA Comments on EPA *Framework for Application of the Toxicity
> Equivalence Methodology for Polychlorinated Dioxins, Furans, and
> Biphenyls in Ecological Risk Assessment* (Draft Framework) (68 *Fed.
> Reg.* 44784 (July 30, 2003)); [Docket ORD-2003-0002]{.underline}

To Whom It May Concern,

The American Forest & Paper Association (AF&PA) is pleased to submit
these comments on the Draft Framework. AF&PA is the national trade
association of the forest, paper, and wood products industry. AF&PA
represents approximately 250 member companies and related trade
associations that grow, harvest, and process wood and wood fiber;
manufacture pulp, paper, and paperboard from both virgin and recycled
fiber; and produce solid wood products.

The Draft Framework provides a discussion of the proposed Toxic
Equivalence (TEQ) methodology for evaluating the ecological risks
associated with complex mixtures of dioxins (PCDDs), furans (PCDFs) and
polychlorinated biphenyls (PCBs). The Draft Framework is a response to
an earlier workshop conducted by EPA and the Department of the Interior
(DOI) on using the methodology in ecological risk assessment. Of
particular note, the report on the workshop indicated that a Framework
be developed that would be useful in "multiple programs ... such as
water quality standards and NPDES/SPDES permits, planning activities
such as Section 303(d) list generation, total maximum daily loads
(TMDLs), waste-load and load allocations, and remediation activities
such as CERCLA cleanups and Natural Resource Damage Assessments (NRDAs)
by natural resource trustees.[^1] The Draft Framework references this
Workshop recommendation in several places (Draft Framework, pages 29,
and 46), and thus it appears that EPA intends the Draft Framework
ultimately to be used for regulatory purposes.

We support EPA's efforts to further the science of ecological risk
assessment in this important area. However, as indicated in the attached
detailed comments which were prepared for AF&PA and the National Council
for Air and Stream Improvement (NCASI), there are many issues which the
Draft Framework would need to address before it could serve its intended
purpose. For instance, several of the Workshop recommendations are not
discussed in the Draft Framework and there are numerous scientific
issues and problems associated with the Framework components. Perhaps
most important, the approach is too uncertain to be used as the basis
for water quality regulatory determinations or as the basis for remedial
activities. EPA should repropose the Draft Framework as a screening
level tool to be used with other lines of evidence, and should not use
the document for regulatory or remedial purposes.

Thank you again for the opportunity to comment on the Draft Framework.
Please feel free to contact John Festa or Jerry Schwartz at (202)
463-2700, if you have any questions about these comments.

> Sincerely,

/s/ /s/

John L. Festa, Ph.D. Jerry Schwartz

Senior Scientist Senior Director,

Water Quality Programs

**[Comments on EPA's Draft]{.smallcaps}**

***[Framework for Application of the TEQ Methodology for]{.smallcaps}***

## [PCDD/Fs and PCBs in Ecological Risk Assessment]{.smallcaps}

Prepared on behalf of the

National Council of the Paper Industry for Air and Stream Improvement

and

The American Forest and Paper Association

by

AMEC Earth and Environmental

October 28, 2003

**[Comments on EPA's Draft]{.smallcaps}**

***[Framework for Application of the TEQ Methodology for]{.smallcaps}***

***[PCDD/Fs and PCBs in Ecological Risk Assessment]{.smallcaps}***

# Introduction

In June 2003, the U.S. Environmental Protection Agency (EPA) released
its 2003 External Review Draft of the *Framework for Application of the
Toxicity Equivalence Methodology for Polychlorinated Dioxins, Furans,
and Biphenyls in Ecological Risk Assessment* (Draft Framework). This
document provides a discussion of the proposed Toxic Equivalence (TEQ)
methodology for evaluating the ecological risks associated with complex
mixtures of dioxins (PCDDs), furans (PCDFs) and polychlorinated
biphenyls (PCBs). In the TEQ method, toxic equivalency factors (TEFs)
and bioaccumulation factors are used to convert concentrations of
individual PCDD/PCDF and PCB congeners, about which limited information
is available, into total concentrations of "dioxin-like" compounds,
which are then evaluated using toxicological information that is known
about 2,3,7,8-tetrachlorodibenzo-*p*-dioxin (TCDD).

The National Council of the Paper Industry for Air and Stream
Improvement (NCASI) and the American Forest and Paper Association (AFPA)
have concerns about the process used to develop the Draft Framework and
believe that there are substantial weaknesses and technical limitations
associated with the use of the proposed approach for ecological risk
assessment. The concerns include the high level of uncertainty created
by combining the substantial uncertainties associated with each step of
the methodology, as well as its emphasis on impacts to individual
organisms rather than population effects.

At the request of NCASI and AFPA, AMEC Earth and Environmental (AMEC)
has prepared these comments on the proposed methodology. These comments
include concerns about the way in which the Draft Framework has been
developed, the technical limitations of the approach itself, and the
difficulties associated with the application of the approach to
regulatory and remedial activities.

######## Development of the Draft Framework

In 1998, EPA and the Department of the Interior (DOI) sponsored the
*Workshop on the Application of 2,3,7,8-TCDD Toxicity Equivalency
Factors to Fish and Wildlife* for the purpose of addressing the
uncertainties associated with the use of this approach in ecological
risk assessment. This workshop, which included representatives from
government, industry, environmental groups, and academia, reached a
number of conclusions regarding the methodology. While some of the
conclusions, which supported the use of the approach, were specifically
cited in the Draft Framework, other conclusions pertaining to the high
levels of uncertainty associated with the approach, which were also
presented in the Workshop proceedings (EPA, 1998), have not been
adequately discussed or addressed in the Draft Framework.

According to the Preface of the Draft Framework, the 1998 Workshop
recommended that further guidance be developed related to the
application of the TEQ methodology. As stated, "\[t\]his framework has
been developed in direct response to that workshop recommendation" but
goes on to state that "it is not intended to be a comprehensive guide to
risk assessment involving dioxin-like compounds." (EPA, 2003; p. iv).
While it is clear that the development of the Draft Framework is an
extension of the recommendations of the 1998 Workshop, the Draft
Framework falls short of the recommendations made by that group. These
inconsistencies are discussed below.

####### [Derivation of the Relative Potency and Toxic Equivalency Factors]{.underline}

The 1998 Workshop participants stated that the development of relative
potency estimate (ReP) and TEF values needed to be "adequately
documented (including specific citations) in order to support the use of
these values in regulatory risk assessment." The workgroup specifically
noted that the World Health Organization panel did not provide adequate
documentation for the selected mammalian TEFs and that this omission was
a "major limitation on the use of the document for risk assessment
purposes" (EPA, 2001; p. 7-8).

These deficiencies have not been adequately acknowledged or addressed in
the Draft Framework. In addition, specific information is not available
to ecological risk assessors concerning the decision-making process that
was used in assigning specific TEF values based on multiple studies.
Although the Draft Framework encourages risk assessors to consider the
appropriateness of the individual TEFs to their site-specific
conditions, without adequate documentation risk assessors cannot
reliably evaluate the appropriateness of the proposed TEFs to
site-specific conditions, species, or endpoints of concern.
Consequently, risk assessors are largely faced with either accepting the
TEFs as presented, without consideration of their appropriateness, or
conducting their own research into toxicological studies to select
alternative values. Although acceptance of the existing TEFs without
consideration of site-specific factors may be inappropriate, the time
and resources associated with having individual risk assessors
independently evaluate the available data makes the alternative approach
unrealistic.

EPA should provide risk assessors with adequate background documentation
for the selection of TEFs so that they can determine whether such TEFs
appear appropriate to their species and endpoints of interest. While the
Draft Framework includes a link to the WHO database of available
studies, this database lists the studies but provides no insight into
the quality of the studies or the way that the studies have been
considered in deriving the TEFs. Thus this database is not particularly
helpful to risk assessors who need to evaluate individual studies for
their quality and/or the species and endpoints of concern. In addition,
because the database is limited in the information that it includes, it
provides no insight into the uncertainties associated with the
toxicological parameters that are derived from them, making it difficult
for risk assessors to evaluate potential uncertainties associated with
their selection, as recommended in the Draft Framework.

[Establishment of Detailed Guidance for ReP/TEF Selection]{.underline}

At the conclusion of the 1998 Workshop, the Planning Group recommended
that a workshop be established to develop detailed guidance on the
selection of appropriate ReP and TEFs. The Draft Framework does not
provide such detailed guidance and does not discuss whether it has been
developed and submitted for peer review. While the Draft Framework does
discuss the selection of alternative ReP and TEF values, relative to
individual species and endpoints of concern, that guidance is vague and
unwieldy in that it leads to many decision points that are highly
subjective with no specific guidance concerning how best to decide among
alternatives.

It is likely that, if risk assessors attempt to select alternative ReP
values, they will be faced with substantial numbers of studies that
focus on species and endpoints that differ from those of concern for a
specific setting. Without clearer guidance as to how to decipher and
select appropriate values based on conflicting data, the process will be
unmanageable, highly subjective, and subject to substantial criticism or
disagreement with reviewers. In this situation, it is likely that risk
assessors will be forced to return to the point estimate TEFs that have
been proposed, thereby undermining the recommended alternative approach.

It is critical that detailed guidance be established for the development
of ReP/TEF values, and that this guidance be peer reviewed prior to the
Draft Framework being finalized. Without such guidance, risk assessors
will likely be faced with blindly accepting the proposed TEF values,
despite the fact that they may not be appropriate for individual
circumstances.

######## Technical Limitations of the Approach

The methodology outlined in the Draft Framework involves the use of both
TEFs and bioaccumulation factors (BAFs) or Biota to Sediment
Accumulation Factors (BSAFs) to convert non-TCDD dioxins, furans, and
PCB congeners into toxic equivalents of TCDD. The approach is highly
controversial within the scientific community, even for the assessment
of human health. Its application to ecological risk assessment is even
more controversial. This is largely due to the fact that the point
estimate TEFs for each taxonomic class have high levels of uncertainty
associated with them because of substantial limitations in the data upon
which they are based. These uncertainties surround the assumptions of
additivity, species sensitivity, dose response, and assignment of point
estimate values that are inherent in the TEF approach. In addition,
derivation of BAFs/BSAFs is greatly complicated by the need to fully
understand how differences among species and aquatic habitats affect the
potential for dioxins, furans, and PCBs to be accumulated in biota.

#### **Lack of Scientific Consensus**

The Draft Framework states "the methodology is well accepted in the
scientific community, in the international risk assessment community,
and within EPA for human health risk assessment" (EPA, 2003; p. 14) and
implies that because of this general "acceptance", it is also
appropriate to apply the methodology to ecological risk assessment.
There is, however, considerable disagreement within the scientific
community about the appropriateness of the approach, even for human
health risk assessment.

The approach has been recommended in EPA's draft Dioxin Reassessment
(EPA, 2000). However, in response to the controversy that has risen in
response to the release of that draft document, Congress has directed
that the National Academy of Science (NAS) should review the draft
Dioxin Reassessment, including its application of the TEQ approach to
PCB congeners. In February 2002, the chairman of a House subcommittee
sent a letter to the EPA Administrator requesting that, due to
"substantial questions regarding the scientific underpinning of the
Reassessment's conclusions about the toxicity of dioxin and
'dioxin-like' compounds," EPA should undertake an agreement with the NAS
to review the draft Dioxin Reassessment, including "the appropriateness
of including 'dioxin-like' chemicals in the risk assessment without
independent empirical review of their effects" (Walsh, 2002).
Thereafter, EPA asked for additional time to convene an Interagency
Working Group (IWG) to review these issues. On February 13, 2003, the
House and Senate agreed to a conference report (House of
Representatives, 2003) on an appropriations bill (later signed by the
President on February 20, 2003 as Public Law 108-7) directing that if
the IWG did not complete its review within 60 days of enactment, EPA was
to contract with the NAS as quickly as possible to review the draft
Dioxin Reassessment, including the issues specified in Chairman Walsh's
letter. Since the IWG did not complete its review by that date, under
Congress's directive the draft Dioxin Reassessment is being submitted to
the NAS for review.

Application of the approach to ecological risk assessment is even more
controversial because of the extremely limited data available for
specific ecological receptors and endpoints, and the enormous
differences among species, in terms of their sensitivity to these
compounds, their metabolic functions, their potential for uptake, and
their susceptibility to specific toxic endpoints. These important issues
need to be addressed before the methodology can be adopted as an
acceptable approach for evaluating the ecological risks of these
compounds.

# Derivation of TEFs

EPA originally developed the TEF approach as an "interim" screening tool
to evaluate dioxins and furans because many congeners lacked specific
toxicity data. Because TCDD had been well-studied and was shown to be
carcinogenic in laboratory animals, a scheme was devised to relate the
toxicity of all other dioxins and furans to the cancer potency of TCDD,
so that a single cancer slope factor could be used for screening-level
human health risk assessments of these compounds. Since that time, the
approach has been expanded to include evaluation of certain PCB
congeners, as well as PCDDs and PCDFs, in the evaluation of ecological
as well as human receptors. In addition, it is assumed that the approach
supports the evaluation of noncancer endpoints as well as carcinogenic
potential.

The approach is based on a number of assumptions that are not well
supported by the scientific evidence and may well not be true. These
include the following: (1) that the toxic effects of all the congeners
in a mixture are additive; (2) that the dose-response curves for all
toxic endpoints and species are parallel with the dose response curve
for TCDD; and (3) that there is no inter-species variability in
sensitivities to these compounds so that all members of entire taxonomic
classes of animals (e.g., fish, birds, mammals) can be represented by a
single set of point estimate TEFs. As discussed below, none of these
assumptions has been established and evidence exists to indicate that
they are likely incorrect.

[Additivity of Individual Congeners]{.underline}

The TEQ approach assumes that the toxicities of all individual congeners
in a mixture are additive. This assumption is not borne out in the
scientific literature and is likely incorrect.

Knowledge of the mechanisms by which AhR-active chemicals cause effects
suggests that the congeners' toxicities represented by TEFs should not
be additive. The AhR binds with a variety of molecules. Whether the AhR
binds with a chemical, and the strength of this binding, is a function
of the shape of the chemical molecule. A chemical that binds weakly to
the AhR may be replaced by a "competitor" chemical that forms a stronger
bond with that receptor so that the binding is competitive rather than
additive.

The fact that a chemical binds with the AhR does not indicate that it
will cause an adverse effect. In fact, chemicals that bind with the AhR
can have a beneficial effect (e.g., triggering a normal physiological
response like enzyme induction), an adverse effect (e.g., overt
toxicity), or no effect. The adverse effects caused by chemicals that
bind with the AhR can range from minor (e.g., inhibiting the production
of certain cells useful in fighting infection) to major (e.g., causing
reproductive disorders). A chemical that binds to the AhR and causes any
effect is called an "agonist." A chemical that binds but has no effect
(or inhibits a "normal" event) is called an "antagonist." The term
"antagonist" results from the fact that chemicals that bind with a
receptor with no adverse effect compete with agonists for sites on
receptors -- while an antagonist occupies the site, an agonist can not
occupy it and cause its effect. Moreover, even agonists can have
antagonistic properties. For example, if an agonist that produces either
a normal physiological effect or a minor adverse effect competes for a
receptor and blocks it from another agonist that causes a more serious
adverse effect, substantial harm has been avoided (Newsted et al., 1995;
Walker et al., 1996). Agonists that have antagonistic properties are
sometimes called "partial" or "weak" agonists.

This understanding of the AhR mechanism substantially weakens the
primary assumption of the TEQ approach that the potencies of individual
agonists can be summed to predict the toxicity of a mixture of agonists
in the body. Where antagonists are present in concentrations higher than
the concentration of agonists, it is difficult for agonists to bind to
receptors. Moreover, partial agonists or incomplete agonists compete
with complete agonists for receptor binding sites. Thus, whenever an
organism contains a mixture of complete agonists, partial agonists, and
antagonists, the total impact cannot possibly be predicted by the sum of
the various agonist concentrations.

Empirical data indicate that some congeners may have antagonistic
properties. For example, Starr (1997) reported that "...some PCDFs
antagonize AhR-mediated responses including fetal cleft palate,
hydronephrosis, immunotoxicity, embryotoxicity and induction of
CYP1A1-dependent activities." Other investigations of mixtures have
indicated that the interactions of congeners are not strictly additive.
In some cases, the observed effects of the mixtures have been lower than
would be expected if the effects were additive (Keys et al., 1986;
Bannister et al., 1987; Haake et al., 1987; Davis and Safe, 1988, 1989,
1990; Biegel et al., 1989; Morrissey et al., 1992; Aarts et al., 1995;
Bosveld et al., 1995; Harper et al., 1995; Harris et al., 1995; Tysklind
et al., 1995; Walker et al., 1996; Zhao et al., 1997a,b). Alternatively,
some studies indicate that the interaction may be synergistic (Bager et
al., 1995; Schmitz et al., 1995; Birgelen et al., 1996) while still
other studies indicate a combination of antagonistic and synergistic
effects, depending on the specific endpoints and species considered
(Silkworth et al., 1984, 1989a,b). Thus, the assumed additivity of these
mixtures does not appear to be demonstrated across congeners and
endpoints in animal studies. In these circumstances, it is unwarranted
to assume that the toxicity of PCDD, PCDF, and PCB mixtures can be
predicted by summing the TEQs for the individual congeners.

Currently the National Toxicology Program (NTP) is conducting research
to test the validity of the additivity assumption. This ongoing research
has not been acknowledged or discussed in the Draft Framework but its
results may be critical for an important assumption upon which the
proposed framework is based. It is important that the methodology
proposed in the Draft Framework not be adopted until NTP has released
its findings that may confirm or refute the key underlying assumption of
the TEQ approach that the toxic effects of TEQ mixtures are additive.

[Shape of the Dose-Response Curve]{.underline}

Because the TEFs are used to equate the toxicity of individual congeners
to that of TCDD (at any dose or concentration), the approach also
necessarily assumes that the dose-response curves for all endpoints and
congeners are parallel to the dose response curve for TCDD. There are
indications, however, that this is not correct (Pohjanvirta et al.,
1995; Putzrath, 1997). In addition, it appears that the doses required
to produce toxic effects vary considerably among congeners and
endpoints.

Safe (1990) evaluated the relative dose response for various dioxin and
furan congeners in terms of the potencies associated with different
endpoints. The relative potencies varied by more than an order of
magnitude depending on the endpoints considered. For example, when
comparing the relative potency of TCDD to that of 2,3,7,8-TCDF at the
EC~50~ level in rats, Safe reported that TCDD was 4 times more potent
for receptor binding, 40 times more potent for thymic atrophy, 64 times
more potent for body weight loss, and 163 times more potent for AHH
induction. Similarly, when comparing TCDD with 1,2,3,4,7,8-HxCDF, the
relative potency for TCDD was higher than the relative potencies for
1,2,3,4,7,8-TCDF by factors ranging from 5 to 73, depending on the
endpoints considered.

[Inter-species Variability]{.underline}

The TEQ approach assigns single compound-specific TEFs for large groups
of organisms (e.g., fish, birds, mammals), without regard for the
variability in sensitivity that occurs among organisms within the same
class. For example, it assumes that the relative level of enzyme
induction seen in a small mammal (e.g., a field mouse) is equivalent to
the relative level of induction that occurs in exposed humans so that
the responses observed in one organism of a certain type is expected to
occur in all organisms in the same class. Thus it makes no adjustment
for variability in species' sensitivities.

Scientific evidence indicates that this assumption is not appropriate.
When looking at different species, Safe (1990) reported differences in
potency for enzyme induction. In summarizing ED~50~ levels for AHH
induction in rat liver, Safe reported that the potency for TCDD was 368
times higher than the potency for 1,2,3,7,8-pentaCDF. However, when the
same endpoint was considered in guinea pigs, the potency of TCDD was
only 21 times higher than the potency of 1,2,3,7,8-pentaCDF. Thus it
appears that there is substantial inter-species variation in sensitivity
of response to PCDD/F congeners, making it inappropriate to assume a
one-size-fits-all estimate of the TEF for all animals within a certain
class.

The Draft Framework acknowledges this limitation in stating that the
"relative sensitivity to dioxin-like toxicity among species that possess
the Ah receptor varies greatly, even within taxonomic class." (EPA,
2003; p. 18). It goes on to state that the "sensitivity of bird species
tested to date to TCDD-induced embryo mortality varies by about
200-fold, with the domestic chicken generally more sensitive than
wildlife species." Given the fact that the majority of TEFs for birds
are based on studies of chickens, these TEFs may overestimate likely
effects in wildlife species to that degree. The Draft Framework also
states that "{t\]wo species that differ widely in their sensitivity to
2,3,7,8-TCDD can have relatively similar RePs for most congeners....In
summary, there are presently insufficient data to determine if there is
any association between sensitivity to 2,3,7,8-TCDD and RePs for
different species." (EPA, 2003; p. 49).

When discussing interspecies differences in sensitivity to TCDD itself,
participants in the 1998 Workshop indicated concern that, while it is
assumed that the effects reported in laboratory studies will result in
population effects in wildlife, this assumption has not been
substantiated with either laboratory studies of wildlife species or
field studies (EPA, 1998). One Workshop participant stated that
"\[t\]his lack of knowledge produces a level of uncertainty that dwarfs
any presented by the TEF/TEQ approach. The current ecological risk
assessment process is seriously compromised by the inability of the
'best available science' to accurately predict effects." (EPA, 1998;
C-E-20).

To address this concern, EPA (2003, p. 20) suggests "risk assessors
should consider the uncertainty introduced when extrapolating from a
species or endpoint for which sensitivity has been established to a
species or endpoint of unknown sensitivity." However, as a practical
matter, if interspecies extrapolation is necessary, it is not likely
that there are adequate data available for the untested species or
endpoint to be able to adequately "consider" the uncertainty, as
recommended. Without any real data, it is impossible for risk assessors
to make an educated guess about the degree or direction of uncertainty
associated with the extrapolation. Thus, such a recommendation, while
appropriate under optimal circumstances, is meaningless in data
deficient situations.

**Derivation of Bioaccumulation Factors**

Because the concentrations of PCDD/Fs and PCBs in abiotic media are not
always measurable or representative of the concentrations found in the
biotic tissues, it is necessary to use bioaccumulation factors (BAFs) or
Biota to Sediment Accumulation Factors (BSAF) to estimate tissue
concentrations of individual compounds before applying the TEQ approach.
EPA acknowledges that BAFs and BSAFs are "the essential connectors of
concentrations of PCDDs, PCDFs, and PCBs in the environment with
concentrations in the diet or relevant tissues of organisms of concern
which are then used to calculated TECs." (EPA, 2003; p. 43). The draft
guidance does not, however, adequately discuss the uncertainties
associated with the derivation of BAFs and BSAFs.

The TEQ approach relies heavily on the assumption that reliable
bioaccumulation factors are available on a site-specific basis or can be
calculated based on data from other locations. The document advocates
the use of chemical-, species-, trophic level- and site-specific BAF and
BASF values (p. 27, Figure 4 and p. 34) and indicates that these data
are needed because of uncertainties associated with extrapolation from
other sites and species due to variation with location, time, and
species, even among species at similar trophic levels.

Despite the acknowledgement of the need to collect site-specific data,
the document concludes that direct extrapolation of BAF/BSAF data from
one location to another can be appropriate. It states that "\[w\]hen the
trophic level, food web, and the sediment-water concentration quotient
are similar for two ecosystems, direct extrapolation of BAFs or BSAFs
from one ecosystem to the other can be accurate if concentrations of
chemicals in water or sediment are defined and measured in a consistent
way for both sites. The Draft Framework goes on to state that "when
conditions are not comparable, as often is the case, BAFs or BSAFs can
be adjusted, using a basic food chain model such as that of Gobas
(1993), for differences in trophic level, food web and \[sediment-water
concentration quotient\]). This will increase accuracy of the BAF or
BSAFs when applied to the unmeasured system. There is a need to acquire
data in case studies in order to validate such extrapolation
approaches." (EPA, 2003; p. 46) There are a number of problems with the
proposed approach.

#### Assumption of Steady State

The Draft Framework states that BAFs and BSAFs are "determined and
applied for conditions that approximate steady-state of the organism
with respect to water and sediments, respectively." (EPA, 2003; p. 34).
While BSAFs and BAFs are assumed to reflect steady state relationships
among contaminant concentration in biota, water and sediments, often
this is not the case. Changes in water and sediment concentration over
time as a result of changing hydrologic conditions or changes in the
concentration of compounds in discharges along with seasonal movements
of fish can all preclude the establishment of a true steady-state. In
river systems where anadramous fish are species of concern, the fish may
never reach steady state with the ecosystem.

#### Site-To-Site Extrapolation

For sites where site-specific data on BSAFs or BAFs are not available,
the Draft Framework recommends that the bioaccumulation factors that
were used for deriving ambient water quality criteria be used. These
values were based on sediment and biota sampling in one Great Lake.
While it was questionably appropriate to apply these factors to the
other Great Lakes waters, they certainly do not provide an adequate
substitute for site-specific bioaccumulation data across the entire
nation. Differences in food web structure, physiological changes during
the season, and feeding locations can result in very different estimates
of uptake even among similar species. For compounds such as PCDD/Fs and
PCBs, for which food web exposures are the driver of overall
accumulation, site-to-site differences in food webs must be accounted
for. Even when site-specific bioaccumulation data are available, the
results are not always consistent. For example, Connelly and Glaser
(1998) reported that even within the same water system, fish at the same
trophic level that both feed on benthic invertebrates had considerably
different bioaccumulation levels.

The transfer of PCDD/F and PCBs from sediments to fish to higher trophic
level organisms is affected by the organic carbon content of the
sediment (or suspended solids), the lipid content and feeding behavior
of the fish, and the feeding range and behavior of the higher trophic
level organisms. Both TOC and fish lipid contents can be highly variable
across different hydrologic systems and within different sections of the
same water system. For example, Currie et al. (2000) reported that the
BSAFs for TCDD in aquatic organisms exposed in a small oligotrophic lake
varied by exposure duration and species. In addition, BSAFs for TCDD
ranged from 0.03 to 0.20 in five predatory freshwater fish species
collected from Lake Ontario (Carey et al., 1990; as cited in Currie et
al., 2000).

The Draft Framework states that site-to-site extrapolation is
appropriate if adjustments are made to reflect differences in lipid
levels between species. There is, however, some indication that lipid
content may not be the only key factor in determining uptake of
PCDD/PCDFs. Miller and Schram (2000) assessed the uptake of
organochlorine compounds by Lake Trout and compared the observed results
to a number of morphometric parameters. Using their total PCB results as
a surrogate for the similarly recalcitrant PCDD/Fs, the authors reported
a positive correlation between total PCB levels in flesh with length
(r^2^=0.69) and age (r^2^=0.71), but no correlation (r^2^=0.03) with
lipid content. This finding is significant and contributes substantial
uncertainty to EPA's approach for site-to-site extrapolation because the
correlation between lipid content and TEQ accumulation is a key
underlying assumption of the proposed approach.

Cook and Burkhard (1998) reported that differences in ecosystem
characteristics (e.g., riverine or lacustrine character, temperature)
and the trophic levels of the fish in those ecosystems are important
considerations for extrapolating BSAFs between different systems. BSAFs
derived for one species in one particular waterbody may not be relevant
to another species in the same waterbody or the same species in another
waterbody. Differences in habitats, life histories, and feeding
behaviors of the fish as well as hydrodynamics of different waterbodies
can substantially affect the potential for certain constituents to
bioaccumulate. Thus the use of surrogate species or waterbodies is
highly uncertain.

#### Use of Food Web Models

While food web models, such as those recommended by EPA when there is
need to extrapolate from one location to another, provide a means of
validating BSAF and BAF data, they depend on adequate characterization
of the site-specific bioaccumulation process in order for them to
produce reliable results. Connelly and Glaser (1998) reported that
within the same waterbody, bioaccumulation in dover sole was two to
three times lower than bioaccumulation measured in white croaker, even
though both belong to the same trophic level and consume benthic
invertebrates. In addition, Connelly and Glaser (1998) attempted to use
food web modeling to predict BSAFs in different sections of the Fox
River study area in Wisconsin. While their model successfully predicted
BSAFs in certain zones of the river, it was unable to do so in other
zones. They concluded that they could not accurately predict
accumulation rates in some other reaches either because the fish were
not in a steady state relationship with the system or the sediment data
did not accurately represent the exposure concentrations to which the
fish were exposed.

# Uncertainties in the Approach

Use of the TEQ approach in remedial and regulatory decision-making
requires the assertion that the selected TEFs are scientifically sound.
There are, however, too many uncertainties associated with the approach
to allow one to conclude that a strong scientific basis exists for its
use. The combination of all of the uncertainties associated with the
selection of TEF values results in a highly unreliable estimate of the
toxic potential for PCBs or PCDD/Fs other than TCDD. When combined with
the additional uncertainties associated with the BAFs/BSAFs the approach
is extremely unreliable and has limited, demonstrated scientific basis.

The Draft Framework acknowledges that the identification,
quantification, and reduction of uncertainties is an important component
to risk assessment. While it repeatedly states that TEFs are no more
uncertain than other uncertainties involved in risk assessment, it
provides no estimates of what the uncertainties may be for other
components of the assessment. In fact, it states that "\[b\]ecause of
the multiple models used for deriving ReP values for a particular
chemical, it is difficult to estimate the variability or uncertainty of
a TEF or RPF point estimate." (EPA, 2003; p. 67). In addition, it
provides no guidance about how the uncertainties involved in this kind
of assessment can be quantified or even estimated.

While the Draft Framework asserts that the uncertainties associated with
the proposed approach are no greater than the uncertainties associated
with the risk assessment process, in general, it does not adequately
address the fact that the uncertainties associated with the TEF
approach, when combined with high levels of uncertainty associated with
other aspects of the risk assessment, may result in wholly unacceptable
levels of uncertainty. Because there is no way to quantify the degree of
uncertainty or to determine its direction, the result is a risk
assessment that is meaningless in terms of making risk management
decisions. The Draft Framework acknowledges this in stating that
"\[s\]ources of inaccuracy and variability have not been adequately
examined experimentally to allow for determination of their relative
magnitudes or their relative contribution to the overall variability of
ReP data used to formulate TEFs or RPFs." (EPA, 2003, p. 67). In
addition, it states that "TEFs and RPFs are point estimates even though
the experimental data from which they are derived may range over several
orders of magnitude. Hence, TEFs and RPFs include uncertainty in the
individual RePs, as well as the uncertainty in the method used to derive
the REF or RPF. Because of the multiple models used for deriving ReP
values for a particular chemical, it is difficult to estimate the
variability or uncertainty of a TEF or RPF point estimate." (EPA, 2003;
p. 67). While the Draft Framework states that TEFs were generally
rounded to the "nearest factor of 5 or 10", this provides little insight
into situations where a point estimate was selected even though the
results of multiple studies varied by orders of magnitude.

There are additional situations where the approach proposed for reducing
uncertainty in the risk estimates actually introduces additional,
unquantifiable uncertainty. For example, the Draft Framework states that
dose metrics derived between the exposure assessment and the toxicity
study should be the same and that if the two dose metrics do not match,
the risk assessor needs to evaluate the direction and magnitude of error
that is introduced by combining mismatched metrics. For example, EPA
acknowledges that the mammalian TEFs are mostly based on administered
doses or dietary levels, rather than tissue concentrations. It goes on
to state that "\[s\]uch relative potency data are subject to variability
associated with toxicokinetic differences between chemicals for
absorption, distribution, metabolism, and elimination." The Draft
Framework suggests that when tissue concentrations are not known, "they
may be estimated from environmental media based on bioaccumulation
factors or models... or bioaccumulation from the diet if dietary intake
and concentrations can be estimated." It fails to recognize, however,
that without specific information on absorption, distribution,
metabolism and elimination, these estimates are merely guesses based on
other species that may or may not have similar metabolic parameters and,
despite their intent of reducing uncertainty to the risk assessment,
actually introduce additional layers of uncertainty.

Uncertainties are multiplicative. When two parameters are each
overestimated by a factor of 10, the result of their combination is an
overestimation by a factor of 100. When an additional overestimation
factor of 10 is introduced, the result is an overestimation of actual
risks by a factor of 1000. Thus, the degree to which uncertainty can be
reduced by avoiding the introduction of additional sources of
uncertainty, the more reliable risk estimates will be.

From a risk management point of view, the real issue is whether the
presence of certain dioxins, furans, and/or PCBs in the environment are
presenting adverse impacts on any communities of ecological receptors
found there. Theoretical calculations of potential risks are meaningless
if the degree of overestimation or underestimation of risk cannot be
identified and considered. The only alternative is to look at community
level impacts directly, rather than modeling theoretical impacts that
are highly uncertain. While the TEQ methodology does provide an
appropriate tool for screening potential risks at a site, it is critical
that the methodology be supplemented with field studies to determine
whether the risks predicted using TEQs are actually occurring in the
wild.

**­Implementation Issues**

As stated earlier, the Draft Framework \"has been developed in direct
response\" to a 1998 EPA/DOI workshop.  EPA indicates in the Draft
Framework that the objective of a risk assessment scenario used at the
workshop was to \"set NPDES permit conditions for multiple sources of
chemicals to an aquatic ecosystem\" including using the TEQ methodology
to set TMDLs.  (EPA, 2003, p. 29, 46).   Several important issues are
associated with the application of the TEQ approach to both risk
assessment and regulatory procedures. These include the need to
field-verify the predicted risks, development of a uniform methodology
for dealing with undetected congeners, development of a methodology for
addressing multiple sources of contamination, and consideration of the
added costs and cost/benefits associated with the approach.

[Verification of the Approach]{.underline}

The Draft Framework fails to identify the fact that the TEQ approach is
a screening level tool that assesses theoretical risks to individuals
but does not provide insight into the potential impacts to site-specific
populations. While the Draft Framework acknowledges that ecological risk
assessment should involve the evaluation of population-level effects, it
does not acknowledge that the proposed TEQ approach only provides a
rough prediction of the potential for individual effects and provides no
insight into the probability that population-level effects will be
elicited. Because the TEQ approach is based largely on cellular and
molecular level responses, it cannot approximate the impact of
site-specific conditions and organism-specific compensatory mechanisms
that may result in no impact or reduced toxicity in the wild.

The population-level approach requires that pertinent population
parameters, such as population size, growth rate, density, and
recruitment, be directly evaluated for the species of interest.
Participants in the 1998 Workshop were especially concerned that the TEQ
approach should not be used instead of field methods to assess impacts
to populations. They expressed a specific concern that "\[a\]nother
factor that may complicate the use of TEFs is the paucity of information
about compensatory mechanisms that may mitigate the effect of
dioxin-like compounds at the population level." (EPA, 1998; p. 21). In
fact, the Workshop Proceedings indicated that "\[s\]ome people felt that
the results of the TEF approach would have to be supplemented with more
information on population dynamics and on the relationship between the
biochemical or molecular endpoints on which the TEFs are based and
effects at the population level before the approach could be used to
decide whether to move forward into a regulatory decisionmaking mode."
(EPA, 1998; p. 45). Workshop participants also stated that "\[w\]hen
excess risk is calculated (e.g., when the summation of TEQ exceeds a
threshold effects value), it is important that the potential for effects
to occur in natural environments (i.e., at the population or community
level) be assessed. There is need to ground-truth the TEF/TEQ approach
such that when this approach is used to demonstrate risk that measured
effects at those exposure levels have been observed in field
populations." (EPA, 1998; p. E-8). The Draft Framework does not,
however, adequately discuss or address the concerns raised by Workshop
participants and instead implies that the TEQ approach is a reliable,
stand-alone predictor of potential ecological risks. In fact, the Draft
Framework appears to minimize the importance of population-level studies
by stating that other lines of evidence "can" be incorporated into the
risk assessment, rather than stating that other lines of evidence
"should" be incorporated (EPA, 2003; p.1).

The absence of a demonstration of the ability of the TEQ approach to
predict population-level impacts underscores the importance of not
relying solely on the TEQ methodology to estimate ecological risks. That
is why it is critical that any estimates of risk using the TEQ approach
be supplemented with field studies to determine whether the predicted
impacts are, in fact, occurring within the ecosystem in question.

# [Consideration of Analytical Detection Limits]{.underline}

The Draft Frameworks states that the "analytical detection levels for
congeners should be lower than concentrations at which important
biological effects may occur." It goes on to conclude that Method 1668
is acceptable for PCBs and the Methods 8290 or 1613 are acceptable for
PCDD/Fs. The Draft Framework states that when analytical detection
limits for individual chemicals are "too large to allow measurement of
concentrations which would significantly add to the TEC", risk assessors
may set concentrations at zero, half the detection limit, or at the
detection limit. It goes on to say that "the best method for handling
non-detect in a particular risk assessment should be determined through
consultation between risk assessors and risk managers early in the risk
assessment process." (EPA, 2003; p. 43). This approach is highly
subjective and can result in vastly different estimates of TEQ,
depending upon the way in which non-detect concentrations are handled.
The Draft Framework should instead make specific recommendations
concerning how to handle non-detect concentrations so that the approach
is applied in a uniform way from site to site.

Another issue to be considered concerns the post-remedial sampling of
media of concern to determine whether post-remedial risks are
acceptable. Despite the use of a high-resolution analytical technology,
EPA approved existing methods (i.e., Method 1613) may not have
sufficient resolution to quantify low levels of TEQ congeners in media
of concern. In addition, the same issues of surrogate species and steady
state, which impact the bioaccumulation impacts, come into play when
attempting to determine whether post-remedial concentrations in water or
sediment are acceptable.

#### 

#### Multiple Sources of Contamination

The TEQ approach can especially present problems when one is attempting
to remediate a site with multiple sources or to derive TMDL or NDPES
permit limits for an aquatic ecosystem with multiple dischargers. 
Examples of each of these scenarios follow. TEQ concentrations measured
in tissues of biota may be the result of uptake of PCDD/PCDFs and/or
PCBs from multiple sources. Because many higher trophic level receptors,
which are the species likely to have the highest levels of
contamination, may have large feeding ranges and may bioaccumulate
congeners from many different sources, it is very difficult to determine
the impact that remediation of one source will have on total
bioaccumulation. For example, consider a site on a river that has
historically released TCDD and that TCDD is detectable in sediments
adjacent to the site. No other PCDD/Fs are detectable in sediments. Fish
collected downstream of the site have non-detectable levels of TCDD but
other PCDD/F congeners are detectable and are high enough that the total
TEQ concentration in fish exceeds the allowable TEQ concentration,
regardless of what concentration the non-detect of TCDD is assumed to
have. Would the sediments with TCDD near the site need to be remediated
in order to reduce the unacceptable level of TEQ in fish, even though
the PCDD/F congeners detected in fish are not from the Site and any
reduction of TCDD concentration in fish would not be measurable? In this
case, removal of sediments containing OCDD will do little to improve the
situation. Clearly the site being evaluated is not responsible for the
unacceptable levels of TEQ in fish.

Similar scenarios are likely to arise in the permitting of discharges to
surface water. For example, what will regulatory agencies do when a
POTW, that has OCDD in its discharge, applies for a renewal of its NPDES
permit and fish in the receiving water have non-detectable levels of
OCDD but have an unacceptable TEQ concentration as a result of an
upstream PCB release? Will the NPDES permit be re-issued to the POTW
with the same OCDD permit limit as before? Will the POTW not be allowed
to discharge OCDD because of the potential to add to the already
unacceptable TEQ concentration in fish? These types of situations are
likely to become much more frequent if PCDD/Fs are to be considered
together with PCBs since many waters may be meeting surface water
quality criteria for PCBs or PCDD/Fs alone, but not when these two
classes of compounds are combined.

[Cost Considerations]{.underline}

The additional analytical costs associated with the use of the TEQ
approach in site characterization can be substantial. For example, the
cost of chemical analysis required to implement the TEQ approach at a
typical PCB site would be up to an order of magnitude higher than the
cost of analysis required to use the total PCB approach. While such
additional costs might be justified if the results of the risk
assessment based on the approach were substantially different and
clearly superior to the results of an assessment based on total PCB
analysis, it does not appear that this is likely to be the case at most
sites. At PCB-dominated sites, the TEQ may have little, if any value,
and will substantially increase the costs of risk assessment and, if
cleanup objectives are expressed in terms of TEQ, the costs associated
with post-remediation confirmation sampling. While such a cost might be
justified if the TEQ approach were a stand-alone methodology (i.e., if
field verification of population effects were not necessary), it is not
clear that the additional costs provide adequate additional benefit when
the methodology must be field verified.

**Conclusions**

The TEQ approach is an appropriate screening methodology for evaluating
potential risks to ecological exposures. However, the method is too
uncertain to provide reliable risk estimates upon which to make risk
management decisions. Instead, it is critical that the potential
estimated risks using the TEQ approach be field-verified to determine
if, in fact, the theoretical risks are representative of actual
site-specific conditions.

On page 6 of the final report on the findings of the 1998 Workshop
Report (EPA, 2001), it is stated that "\[a\]s with any method,
appropriate caution should be exercised to avoid misuse or application
of the methodology to situations where the underlying assumptions are
known not to be valid." As discussed in this White Paper, it is clear
that many of the assumptions underlying the TEQ methodology are not
valid. The underlying assumptions of the methodology, including the
additivity of toxic effects, parallelism of dose response curves, and
similar sensitivity among species have all been demonstrated, in the
scientific literature, to be questionable.

In addition, for the BAFs/BSAFs that are necessary for implementation of
the approach to be useful, there must be some indication that biota are
at steady-state with their surroundings, and that water/sediment
sampling is representative of the locations at which the biota are
exposed. If BAFs/BSAFs from other waterbodies are used, the methodology
also needs to demonstrate that the extrapolated estimates of uptake of
different congeners in the source water body are applicable to the water
body of interest. A summary and comparison of available BAFs/BSAFs
derived for different waterbodies would provide important insight into
the potential uncertainty of applying the bioaccumulation portion of the
proposed methodology to ecological risk assessment.

While the TEQ approach proposed in the Draft Framework document may
provide a reasonable screening level tool to conduct initial risk
evaluations of ecological receptors, the approach itself is too
uncertain to provide a basis for remedial or regulatory decisions.
Uncertainties about the ability of molecular level laboratory studies to
predict adverse effects in the wild, where compensatory mechanisms may
come into play, and the ability of modeled bioaccumulation factors to
predict actual uptake, make it necessary to evaluate ecosystems on a
population level to determine if adverse effects are in fact occurring.
We recommend that the TEQ methodology be re-proposed as a screening
level approach that, when considered with other lines of evidence
concerning population level effects, can provide insight into the
potential ecological effects posed by exposure to PCDDs, PCDFs and PCBs
at a site.

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