Document ID: EPA-HQ-ORD-2003-0002-0025
Agency: epa
Document Type: Supporting & Related Material
Title: 
Posted Date: 2008-06-16T04:00Z

Response to Comments in the External Peer Review Report on the   SEQ
CHAPTER \h \r 1   SEQ CHAPTER \h \r 1 Draft Framework for Application of
the Toxicity Equivalence Methodology for Polychlorinated Dioxins,
Furans, and Biphenyls in Ecological Risk Assessment

The purpose of this summary is to provide a disposition on EPA’s
response to major comments raised as part of the External Peer Review of
the Agency’s draft “Framework for Application of the Toxicity
Equivalence Methodology for Polychlorinated Dioxins, Furans, and
Biphenyls in Ecological Risk Assessment” (the Framework).

The peer review offered a number of general comments, as well as more
narrowly focused and editorial comments from individual reviewers.  Many
of the comments from individual reviewers have been incorporated into
the Framework and are not presented here.  The peer review was conducted
over several months from October 2003 to February 2004, culminating with
a conference call meeting on December 5, 2004 and final report on
February 9, 2004.  External peer reviewers included:

	Dr. William J. Adams, Rio Tinto, Magna, UT

	Dr. Scott B. Brown, Environment Canada, Burlington, Ontario

	Dr. Peter L. deFur, Environmental Stewardship Concepts, Richmond, VA

	Dr. John P. Giesy, Michigan State University, East Lansing, MI

	Dr. Mark E. Hahn, Woods Hole Oceanographic Institute, Woods Hole, MA

	Dr. Barbara L. Harper, AESE Inc., West Richland, WV

	Dr. Bruce K. Hope, Oregon Department of Environmental Quality,
Portland, OR

	Dr. Sean W. Kennedy, Environment Canada, Ottawa, Ontario

	Dr. Charles A. Menzie [Chair], Menzie-Cura & Associates, Inc.,
Winchester, MA

	Dr. Christopher D. Metcalfe, Trent University, Peterborough, Ontario

	Dr. Richard E. Peterson, University of Wisconsin, Madison, WI

	Dr. Martin Van den Berg, Utrecht University, Utrecht, Netherlands

EPA appreciates the useful feedback and recommendations of the peer
review panel and believes their comments have further improved the
overall quality of the Framework.  A summary of the major comments and
EPA’s responses are organized as follows: 

Key issues identified in the reviews 

Overview of general comments 

Responses to charge questions 

Overall Summary:

Comment:  There was broad agreement that the document met its major
goals and objectives. Therefore, our comments are intended to provide
helpful feedback on how the Framework can be made more useful to the
intended audiences. Often, this involves clarification. In some cases,
our comments reflect areas where one or more of us disagree with
technical statements made in the Framework document. These technical
issues were discussed during our December 5th conference call. For the
most part, the tenor of our comments is captured by a response (caveat)
from John Giesy:

In general, this document is very useful and a much needed improvement
on previously available documents and guidance… there are many very
positive aspects to the document, but to be concise, I will limit my
comments to those where I think that the document can be improved.  If I
am silent on an issue or section of the document it indicates my
concurrence with those conclusions or guidance.

Key Issues Identified in the Reviews

The following issues were identified prior to our December 5th
conference call and considered during the call. There was general
agreement that these were the major issues with respect to our review. 

Management-related considerations including how to judge the strengths
and limitations (and costs) of the Toxicity Equivalence Methodology
relative to other approaches

Clarification of text and consistent use of terminology

Approaches to estimating the bioaccumulation of chlorinated compounds in
animal tissues 

More detailed information on dose response

Quantification of uncertainties and possible use of probabilistic
methods

EPA Response: The key issues identified in section A are a subset of the
more detailed comments in Sections B and C, and are therefore addressed
in the response to the more specific comments.

Overview of General Comments

Provide a short section or perhaps only a few paragraphs in either an
Executive Summary or in the Introduction that gives the reader a more
complete view of the pluses and minuses of the method. This was
considered particularly important for risk assessors and managers who
are making decisions on how to proceed for particular sites. It would be
useful to a greater range of practitioners, and in a more equitable
fashion, if the document directly addressed both the benefits and the
costs (burdens) of its key technical recommendations.  No matter how
“technical” a guidance document tries to be, its recommendations
will have non-technical implications, such as driving increased costs
onto regulated parties or greater review burdens onto regulators.  Hope,
Harper, and Menzie provide further comments concerning this.

EPA Response:  Discussion of the prerequisites, strengths, and
limitations to be considered in applying the toxicity equivalence factor
(TEF) methodology is provided in Sections 1, 2, 3.1, 3.1.1, and 3.2.2 of
the Framework.  Methodological considerations associated with using the
TEF methodology are presented in Sections 3.1 and 3.2 to allow those
conducting risk assessments for dioxin-like chemicals to consider the
strengths and limitations associated with using the TEF methodology
against other methods they may be considering.  The type and number of
other approaches that could be considered will be specific to each
ecological risk assessment, such that comparisons of costs and benefits
is best conducted during the planning and problem formulation phases of
the specific ecological risk assessment.  Additional discussion in this
regard has been added to the Framework.  Specifically, Section 3.1
(Considerations in Planning) raises the issue that costs and benefits
need to be considered during the planning phase of an ecological risk
assessment.  However, since costs and benefits will vary depending on
the scope and objectives of a specific ecological risk assessment and
will also vary over time, EPA does not believe it is appropriate to
provide a specific comparative analysis within the Framework.  As
appropriate, EPA Offices and Regions may consider costs (monetary as
well as other resource requirements) as they implement the Framework for
their individual programs.  

Provide an illustrative example(s) for the application of the method.
This example(s) could be placed in an Appendix. Most reviewers felt that
this would help people less familiar with the process to follow the
methodology. The Group debated whether this example(s) should use real
numbers or simply illustrate the process. There was a concern that the
use of numbers would lead readers to view the numbers (e.g., for BSAFs)
as the ones that would be used for other sites. The reviewers felt that
any use of examples should be caveated to make sure the readers were
aware that these were intended only for illustration. There was some
discussion on using sensitivity analysis to show how alternative
decisions can influence the outcomes of the assessment. Sensitivity
analysis could also be used to help judge which parts of the assessment
contribute the most uncertainty. This discussion led to either including
an example or including some discussion of the value of sensitivity
analysis (perhaps in the Uncertainty section).

EPA Response:  EPA has integrated illustrative examples in the Framework
within the sections discussing the individual steps in applying the TEF
methodology.  Application of the TEF methodology is illustrated through
the examples provided in Tables 4, 5, and 6, which show how to calculate
toxicity equivalence concentrations (TECs) in fish tissues, bird
tissues, and a mammalian (otter) diet, respectively.  A note has been
added to Tables 4, 5, and 6 to make clear that the data are for
illustrative purposes only and are not recommended default values. 
Application of the selection logic laid out in the relative potency
matrix (Figure 11) is illustrated through the examples provided in
Tables 7 and 8, which show how to array and select relative potency data
for birds and mammals, respectively.  

In lieu of providing another case study, the Framework references the
Workshop Report on the Application of 2,3,7,8-TCDD Toxicity Equivalence
Factors to Fish and Wildlife (U.S. EPA, 2001a) that includes case
studies wherein the TEF methodology is applied to two different
ecosystems.  The workshop report is available on-line
(http://cfpub.epa.gov/ncea/raf/recordisplay.cfm?deid=23763) and will be
linked to from the same web site on which the final Framework is posted.
 In addition, references to a recent peer-reviewed publication (Burkhard
et al., 2006) that describes the basis for and examples of extrapolating
bioaccumulation factors (BAFs) and/or biota-sediment accumulation
factors (BSAFs) across ecosystems have been added in appropriate places
in the Framework.  

The document makes various references to other methods for doing PCB
risk assessment (specifically aroclors and homologues [totals]). In
general, the document points out the advantages of the congener approach
relative to these other approaches. However, the document also notes
that there are ecological receptors and toxicological endpoints that
cannot be addressed with the TEF/TEQ approach (e.g., bottom of p. 5 and
top of p. 11). This leaves open a question on how to best approach sites
contaminated by a broad spectrum of PCBs. The document should provide
clarification on this so that risk assessors can have a better
understanding of how to use the TEF/TEQ approach in concert with other
approaches for assessing risks associated with PCBs. 

EPA Response:  Section 2.1 has been revised to clarify that the
Framework applies only to dioxin-like chemicals, and hence, only
dioxin-like polychlorinated biphenyls (PCBs).  The non-dioxin-like PCB
discussion highlights that these PCBs may cause toxicity via mechanisms
independent of the aryl hydrocarbon receptor (AHR), and because the TEF
methodology will only account for dioxin-like activity of PCBs,
non-dioxin-like PCBs would need to be assessed using another
approach/analysis, just like any other chemical of concern that may
co-exist with the dioxin-like PCBs.  The Framework is not a
comprehensive guide to conducting a risk assessment for dioxin-like or
non-dioxin-like chemicals; hence, to elaborate further on the
appropriate analysis for addressing risks of non-dioxin-like chemicals
is outside the scope of the Framework.  

Does the TEF/TEQ methodology require measurement or estimation of all
dioxin-like compounds including dioxins, furans, and PCBs in order for
it to be valid? There are numerous investigations underway in which PCBs
are being analyzed on a congener-specific basis but where analyses are
not being carried out for dioxins and furans. This is fairly typical for
a site where PCBs are considered the main issue. Inclusion of
chlorinated dioxins and furans can be accommodated but at a significant
additional analytical cost. The document should be clear on this matter
one way or the other and should include some discussion of the
limitations (i.e., uncertainties) of including only PCBs in the
approach.

EPA Response: The TEF methodology is a tool that facilitates cumulative
assessment of any “dioxin-like” chemical, i.e. any that acts via the
AHR.  The “validity” or robustness of a risk assessment could be
influenced by whether all chemicals acting via the AHR have been
cumulatively accounted for and whether only AHR-agonists are accounted
for in the TEC; the TEF methodology allows one to include any or all
AHR-receptor agonists.  The chemicals of concern in any ecological risk
assessment are site or assessment specific.  Therefore, the decision
regarding what chemical(s) should be included in the analysis plan for
an ecological risk assessment should be discussed and decided upon
during the planning and problem formulation phases of the ecological
risk assessment.

Figure 6 can be modified (or additional figures generated) to illustrate
for the reader the specific characteristics of dose response curves for
fish, birds and mammals. This would make it so much easier for the
ecological risk assessor who is uninitiated in the use of the toxicity
equivalence methodology to grasp the concept. Thus, the addition of TCDD
dose response curves for a sensitive, population-relevant endpoint for a
representative fish, bird and mammal would be valuable additions. It
would be helpful to designate, for teaching TEF methodology only, a
“hypothetical” threshold or action level for TCDD for each species
to which the calculated TECs could be compared.

EPA Response:  A single, generic dose-response is provided in Figure 7
of the revised Framework.  It is representative of the type of
dose-response that is typical of fish, birds, and mammals for an
endpoint that is sensitive and highly relevant in ecological risk
assessment (i.e. early life stage mortality).  Definition of a threshold
or action level is an assessment-specific activity and is therefore not
included in the generic figure.

EPA believes a generic illustration is appropriate for the Framework
because the document is not intended to be a comprehensive guide to risk
assessment of dioxin-like chemicals. In addition, the intended audience
for the Framework is risk assessors who have a working knowledge of
EPA’s Guidelines for Ecological Risk Assessment (U.S. EPA, 1998) and
are familiar with issues related to conducting risk assessments for
dioxin-like chemicals.  EPA has added text to the Preface and the
Introduction clarifying the scope and intended audience for the
Framework.

The document should be a little more critical of the existing WHO
values. This might be handled with a text box in the Introduction. In
this respect it should be mentioned that the eco-TEFs determined by WHO
for fish and birds have often been determined with a minimum available
data set. As such, this limitation certainly represents the observed
difference between birds or fish versus mammals in TEFs. However, it
should be realized that at the time no better choice could be made due
to the limited information available. Thus, the eco-TEFs derived in 1997
should be considered as interim and preliminary values that definitely
do not have the accuracy and detailed information that has been used for
establishing the mammalian TEFs. The EPA should allow itself more to
express this higher uncertainty in bird and fish TEFs where appropriate.
Furthermore it could also be suggested that the database should be
expanded and the 1997 WHO eco-TEFs being reviewed within the near future
to obtain a higher degree of certainty. Such a revision would likely be
done within an international framework such as WHO-IPCS.

EPA Response:  EPA’s position that it is appropriate to use the World
Health Organization (WHO) class-specific TEFs in ecological risk
assessment is supported by the conclusions from two World Health
Organization expert meetings (Van den Berg et al., 1998; 2006), an
EPA-DOI expert workshop (U.S. EPA 2001a), and the recent National
Research Council (NRC) report (2006).  In each of these reports, the
experts convened agreed that application of TEFs for the purpose of risk
assessment is currently an appropriate and scientifically defensible
approach.  Furthermore, the participants in the EPA-DOI expert workshop
on the application of the TEF Methodology in ecological risk assessment
concluded that the uncertainty in the TEF approach is not greater than
the overall uncertainty of the ecological risk assessment process (U.S.
EPA, 2001a).

The Framework acknowledges that when applying the TEF methodology, one
expected approach is to use the TEFs adopted by the WHO in 1998 and 2006
(TEF-WHO98/05 values).  However, a large proportion of the Framework
(Chapter 3) is dedicated to describing a logical way in which risk
assessors can organize relative potency data (including the WHO TEFs)
according to species similarity, endpoint relevance, and dose relevance
and consistency (Figure 10 and Tables 6 – 8) in order to understand
the strengths, limitations, and uncertainties associated with such data
and select the relative potency values that are most appropriate for a
particular ecological risk assessment.  

The table of BSAFs generated much discussion concerning the source(s) of
these values as well as concerns that they might be viewed as default
values. There was strong sentiment that they should not be portrayed as
default values. The reviewers felt that the legend should be expanded to
make that clear and that information should be given on where these
values did come from.

EPA Response:  All the values used as BSAFs, BAFs, or BCFs are
illustrative examples.  A note has been added to Tables 4, 5, and 6 to
clarify that the data are for illustrative purposes only and are not
recommended default values.  

Responses to Charge Questions

 

Perhaps include a few sentences about risk methods that could be used in
addition to the hazard quotient method. Examples are probabilistic and
joint probability analysis.

EPA Response:  Section 3.4.1 has been revised to acknowledge that the
hazard quotient is but one simple risk estimation method and that other
methods for risk estimation may be used in an ERA.   However, because
the Framework is not meant to be a comprehensive guidance for conducting
an ecological risk assessment for dioxin-like chemicals, the reader is
referred to EPA’s Guidelines for Ecological Risk Assessment (U.S. EPA,
1998) for more information on additional methods.

Move the "Conclusions" upfront and make it into an Executive Summary or
part of the Introduction. Organizing the document in this manner will
enable first-time readers to obtain an overview of the methodology, and
important considerations associated with it, before they enter the
detailed portion of the guidance.

EPA Response: EPA has retained the Conclusions section at the end of the
Framework.  However, broad conclusions have also been incorporated into
the Preface and Introduction.

Check all figures and tables for complete legends so that the figure or
table can stand alone.  The text could make more use of the figures and
tables, so it is worth another read to be sure that nothing has been
overlooked.

EPA Response: Figure legends and Table titles have been modified as
suggested.  

A casual reference is made on P. 20 (Line 20) to the use of
“uncertainty factors”. These are often used for interspecies
extrapolations. However, this is the only place the matter is discussed.
Is this Framework suggesting the use of interspecies extrapolation
factors for developing TECs? If so, that is an important aspect of the
method. Either develop that a bit further or do not raise the issue only
in this casual way.

EPA Response: EPA is not suggesting the use of uncertainty factors for
developing TECs.  The noted reference to “uncertainty factors” has
been removed to prevent confusion. 

Consider breaking out Table 3 to provide information for each major
animal class. Each table would provide information on different species
of fish, birds, and mammals. The reason for breaking Table 3 into three
tables is that early life stage toxicity is a very relevant endpoint for
ecological risk, yet the “profile of TCDD effects” that characterize
early life stage toxicity in fish and birds, respectively, is not
clearly illustrated in Table 3 or anywhere else in the document.  The
adverse developmental effects caused by exposure to TCDD in for example
egg laying fish and related AhR agonists (edema, impaired jaw
development, impaired heart development and function, reduced trunk
blood flow, anemia, growth retardation, and mortality) needs to be
captured in the mind of the reader of this document (along with the well
known effects on enzyme induction).  Table 3 simply does not accomplish
this objective. If possible, it would help to give the reader a feel for
the relative sensitivity of the endpoints. This might be done with a
“+” to “+++” type approach.

EPA Response:  Table 3 has been modified to provide information on each
of the major animal classes as a group for endpoints that are unique or
particularly relevant for characterizing dioxin-like toxicity.  The
broad endpoint of developmental and reproductive toxicity is now
included in the table.  However, detail about specific species and
magnitude of effects is not included given that the Framework is not
intended to be a comprehensive guide for conducting the effects
characterization for an ecological risk assessment.  The Table does
include extensive references to guide risk assessors in their further
investigation of the literature pertinent to conducting an effects
characterization.  In addition, EPA has previously published
comprehensive reviews of available toxicity data for dioxin-like
chemicals that are referenced in the document (U.S. EPA 1993).

Some qualification is needed in connection with information presented on
mono-ortho PCBs (in particular consider the use of “less than”
indicators as was originally provided by WHO). Specifically, it was
noted that underlying research indicates that mono-ortho PCBs are not
toxic to fish.  Use of the upper range of the TEFs (i.e., 0.000005) for
the mono-ortho substituted PCB congeners in fish will overestimate the
TEC. At some points in 3.2.1.1 it might be useful to expand a bit more
in the basic difference between the species sensitivity for dioxin like
compounds and the relative potency differences e.g. observed between
mammals and fish for MO-PCBs.  It should be emphasized that in the
future, risk assessment should more be based on internal
dose/concentrations levels than administered dose/uptake is essential to
obtain more information regarding differences in species sensitivity for
AhR mediated mechanism.

EPA Response:  The “less than” indicators associated with the WHO
TEFs for fish for mono-ortho PCBs are present in Tables 2 and 4. 
Furthermore, a paragraph comparing the sensitivity of fish, birds, and
mammals to mono-ortho PCBs is included in Section 3.2.1.1.  Language
describing how internal dose/concentration reduces the variability in
toxic effects thresholds and the need for more approaches based on
internal dose/concentration data has also been added to Section 3.2.1.1.

Section 3.3.1.3 discusses choices for exposure dose metric. It would be
helpful to emphasize the importance of insuring a proper match of dose
to effects as part of Planning. Look especially at the last paragraph on
p. 32.

EPA Response:  This discussion has been highlighted by repeating it in
Section 3.1 Considerations in Planning.

Section 3.3.2.1 could be set up better. It needs a better introduction.
Consider moving the second paragraph (P. 48 Line 11) to after the
current third paragraph (at Line 29).

EPA Response:  Much of Section 3.3.2.1 was redundant with other parts of
the document. Therefore Section 3.3.2 has been extensively revised.  

Section 3.4.2 needs a conclusion. It also has embedded within it various
screening tests. Because these are not recommended as lines of evidence
for risk characterization, do these belong in this section? Should these
types of tests be given their own section, perhaps in an early tier
where screening may be appropriate?

EPA Response:  A conclusion has been added to Section 3.4.2. The experts
at the EPA/DOI workshop concluded that although screening bioanalytical
tools should not be used as an alternative to congener-specific analysis
and the TEF methodology, they may be considered as additional lines of
evidence in a risk characterization.  Thus, EPA believes that such tests
should be discussed in Section 3.4.2 (Lines of Evidence).

On P. 68, Lines 3 – 5, a method is suggested involving the use of
ranges of RePs. Is this appropriate for this document? If there is a
desire to evaluate uncertainties, perhaps an explicit discussion should
be put together on how to quantify this.

EPA Response:  EPA believes that it is appropriate to use alternative
ReP values, when available, to describe the range of possible risk
values.  This approach is consistent with EPA’s Guidelines for
Ecological Risk Assessment and with the Office of Management and
Budget’s (OMB) 2007 memo on Updated Principles for Risk Analysis
(http://www.whitehouse.gov/omb/memoranda/fy2007/m07-24.pdf).  

 

Why use the term ReP to represent Relative Potency?  One should be able
to represent two words with two letters (RP).

EPA Response:  The ReP terminology, definition, and acronym were adopted
directly from those agreed to at two World Health Organization
international consultations (Van den Berg et al., 1998; 2006).

There was strong sentiment that the acronym (term) TEQ should be
retained rather than TEC. The term ‘TEQ” is so well entrenched in
the literature that introducing the new term “TEC” would only add to
the confusion.  

EPA Response:  EPA chose to retain the acronym TEC to represent TCDD (or
Toxicity) Equivalent Concentration because it more accurately represents
the fact that the end product of applying the TEF methodology is a
concentration, and it is more consistent with the construction of the
acronyms derived for other terms associated with the TEF Methodology as
established by the World Health Organization (Van den Berg et al., 1998;
2006).  EPA reviewed the responses of each of the peer reviewers
regarding this charge question.  Six of the reviewers provided positive
comments on the terminology section without reservations regarding the
introduction of the term TEC to represent TCDD (or Toxicity) Equivalent
Concentration.  Four reviewers did not address the issue specifically. 
Only two of the twelve reviewers expressed concerns about introducing
the term TEC.  

Analogous acronyms to TEF have also been REP, RPF and RP. It was
suggested that REP, RPF and RP be added in the table as analogous
acronyms. 

EPA Response:  These terms have been added to Text Box 1 in the revised
Framework.

Consider moving definitions on p. 4 to the beginning of 1.1. For a
Framework document, it is most useful to present the definitions and
then follow with the rationale for what is being proposed. 

EPA Response: The suggested change had been made in the revised
Framework.

Inconsistent use of other terminology currently in the document can lead
to confusion. To avoid this EPA should consider having the document
reviewed by people less familiar with the methodology. Members of the
peer review group identified the following terminology issues and have
suggested changes: 

Readers of this document will find it confusing that the words:
compound, chemical, and congener are used interchangeably.  This is
especially problematic when TEFs are listed for “congeners” and the
type of chemical analysis required to measure exposure to PCDDs, PCDFs,
and PCBs is referred to as being “congener-specific”. It is
suggested that the phrase “dioxin-like congener” or “dioxin-like
compound” be used to insure clarity.

EPA Response: For consistency EPA chose to use the term dioxin-like
chemical(s) when referring to the whole group of PCDDs, PCDFs, and PCBs
throughout the document.

The symbol (IIsocw) used to describe the sediment-water concentration
quotient appears unconventional.  The use of the II in the symbol is not
intuitive.  Various symbols have been used to describe sediment water
partitioning such as Kd or Kp or K 

∏socw is well established in the peer-reviewed literature (e.g.,
Burkhard, 1998, 2003; Burkhard et al., 2003a; 2008) and has been defined
and used in previously reviewed and published EPA documents (U.S. EPA
1995a, 2000, 2003b).

Consider how confusion around the word “receptor” can be reduced.
The term “receptor” is used both to refer to the aryl hydrocarbon
receptor (AHR) and to “ecological receptors” (meaning target
species, e.g. p.14, 28).  The term “receptor” has a specific meaning
in pharmacology, defined more than 100 years ago, and its use in
reference to the AHR is consistent with that.  Using “receptor” in
the context of a target species, while common in ecological risk
assessment, is potentially confusing.  The term “target” or
“target species” would be more descriptive and less ambiguous.

EPA Response:  The term “receptor” is used exclusively to refer to
the aryl hydrocarbon receptor (AHR) in the revised Framework. 
“Ecological receptors” are simply referred to as species.

Similarly, the term “stressor” is often used in the document in
reference to the chemicals that act through the AHR (e.g.
“AhR-mediated stressors”, p. 14, 28).  Why not simply say
“chemicals”?  (Note also that the chemicals are not AhR-mediated,
their effects are.) 

EPA Response:  The term “ecological stressor” has been changed to
“ecological endpoint,” and “stressor” has been changed to
“chemical stressor.”  The terminology “stressor-response
profile” has been retained, since this term is consistent with EPA’s
Guidelines for Ecological Risk Assessment (U.S. EPA, 1998).

P. 1, Line 10. Add after the sentence ending with “situations.”
“In this document, the term “dioxin-like effects” and
“dioxin-like compounds” are used to refer to those effects that are
similar to those caused by 2,3,7,8-TCDD and for those compounds that
exert such effects through binding with the Ah Receptor.

EPA Response:  The Framework has been revised consistent with this
suggestion.

The term “potency” should not be used as a stand alone word at any
place in this document.  The potency of every dioxin-like congener
should always be mentioned relative to 2,3,7,8-TCDD as relative potency.
 In the vast majority of the framework document relative potency is
used.  However, there are a few places where “potency” only is used
and where this occurs it needs to be corrected.   The same comment
applies to the use of “potency factor” in place of the correct term,
“relative potency factor”.

EPA Response: The suggested change has been made in the revised
Framework.

 

Provide a brief comparative discussion of the alternative methods. This
might involve the preparation of a sub-section entitled “Advantages
and Limitations for the TEQ Methodology” This might be placed in the
Introduction. Two reviewers suggested giving an actual example that
compared the methods (e.g., total vs. TEQ vs Aroclor). This would serve
to show how uncertainty is reduced through using the TEQ methodology.
For amplification see comments of Hope, Hahn, and Menzie.

EPA Response:  Discussion of the prerequisites, strengths, and
limitations to be considered in applying the TEF methodology is provided
in Sections 1, 2, 3.1, 3.1.1, and 3.2.2 of the Framework. 
Methodological considerations associated with using the TEF methodology
are presented in Sections 3.1 and 3.2 to allow those conducting risk
assessments for dioxin-like chemicals to consider the strengths and
limitations associated with using the TEF methodology against other
methods they may be considering.  The type and number of other
approaches that could be considered for PCBs as well as other
dioxin-like chemicals will be specific to each ecological risk
assessment, such that comparisons of strengths and limitations is best
conducted during the planning and problem formulation phases of the
specific ecological risk assessment.  Since strengths and limitations
will vary depending on the alternatives available and the scope and
objectives of a specific ecological risk assessment, EPA does not
believe it is feasible to provide a specific comparative analysis within
the Framework.  However, EPA has inserted several references to a recent
document addressing the benefits of PCB congener analysis (U.S. EPA,
2005).  As appropriate, EPA Offices and Regions may consider strengths
and limitations of all methods deemed available or feasible as they
implement the Framework for their individual programs.  

Point out that the method is applicable to vertebrates but not for
invertebrates. Note that there are non-dioxin-like effects that can be
important for invertebrates and that may need to be evaluated using a
separate methodology. Consider changing the title of this document to
reflect that the TEF/TEQ methods applies to fish and wildlife (to
distinguish it from what might be needed for invertebrates.) See Adams
comment on Daphnia.

EPA Response: Section 3.2.1.1 has been revised to include a paragraph
addressing the insensitivity of invertebrates to dioxin-like chemicals
and the recent findings that AHR analogs found in some invertebrates are
unable to bind prototypical AHR agonists, thus, providing a mechanistic
understanding of the relative lack of sensitivity in invertebrates. 
This paragraph also addresses the Adams comment on Daphnia. 

The identification and functional characterization of AHR in a variety
of species is an active area of research (Hahn et al., 2002a, b; Jensen
and Hahn, 2001; Yasui et al. 2004, 2007).  The suggested title change
would exclude the applicability of the Framework to a whole class of
organisms if the current understanding of invertebrate sensitivity to
dioxin-like toxicity were to change as a result of on-going research. 
Therefore, EPA has not made the suggested change to the title of the
Framework.  

 

Consider providing a bit more guidance relative to the development of
tissue concentrations estimated from sediment or dietary exposure.  In
those cases, it is imperative to consider the trophic transfer and
biomagnification that occurs from fish to bird species.  The use of a
model such as that proposed by Gobas (1993) should not be thought to be
optional.

EPA Response:  The Characterization of Exposure section of the Framework
(Section 3.3.1) has been reorganized and revised extensively.  In the
revised Framework, Sections 3.3.1.4 and 3.3.1.5 are dedicated to
providing guidance and illustrative examples of how to derive tissue
concentrations in fish (Table 4), tissue concentrations in bird (Table
5) and mammalian dietary concentrations from sediments using BSAFs
(Table 6).  The bird example (Table 5) illustrates how to calculate
tissue concentrations in bird eggs, rather than through adult bird diet,
because early life-stage toxicity is more relevant assessment endpoint
for dioxin-like toxicity, and available TEFs for birds are largely
tissue-based rather than based on dietary intake.  Additional discussion
has been added to Section 3.3.1.5, and questions have been added in Text
Box 6 regarding how to consider biomagnification and extrapolation of
BAF or BSAFs across sites using food-web models (e.g. Gobas, 1993 and
Gobas et al., 1998).  In addition, a reference (Burkhard et al., 2006)
to a recent demonstration of a “hybrid modeling approach” using
BAFs/BSAFs and food-web modeling to account for trophic transfer and
biomagnifications has been added.

Comment C.4.2:  The document should address the issue of non-detects.
Consider developing a short section for the main portion of the document
or, alternatively, treat this in the uncertainty section. Several
reviewers felt this is an important issue with regard to the low levels
of congeners that occur in some media. A source of uncertainty is the
change in detection levels from one study to the next or at different
times in the same study. (See de Fur and Giesy for further discussion.)

EPA Response:  The issues of analytical methods and detection limits are
overall risk assessment issues, not issues specific to the TEF
methodology.  In addition, the analytical methods and detection limits
issues are specific to each ecological risk assessment and are therefore
best addressed during the planning and problem formulation phases of the
specific ecological risk assessment.  While the TEF methodology does not
dictate what analytical method or detection limits need to be used,
methodological considerations associated with using the TEF methodology
are presented in Sections 3.1 and 3.2.  In section 3.3.1.1, it is
explained that the best method for handling non-detects in a particular
risk assessment should be determined during planning and/or problem
formulations phase(s) of the risk assessment.  In section 3.4.3.2.2,
uncertainties associated with characterization of exposure, including
detection limits, are discussed.  In both Sections 3.3.1.1 and 3.4.3.2.2
a reference to other EPA guidance that addresses this issue is provided.

One reviewer expressed concern about applying TECs in the diet.  This
concern is based in part on the fact that each congener not only has its
own unique ReP or TEF, but also a unique BAF.  Thus, the use of TECs in
dietary items could lead to additional variability in the analysis. 
However, as long as the dietary item is not predicted the use of TECs in
dietary items is appropriate.  More discussions of the limitations of
this use of TECs would be useful. This comment was not discussed further
during our phone conversation.

EPA Response: The Characterization of Exposure section of the Framework
(Section 3.3.1) has been reorganized and revised extensively.  In the
examples presented in Section 3.3.1.5 (and in equations 3-3 and 3-4), it
is clear that bioaccmulation factors (BAFs or BSAFs) are
congener-specific, that is, unique for each congener.  EPA believes the
least amount of variability and uncertainty in the TEC is achieved by
using a congener-specific TEF and B(S)AF in calculating the TEC.  

In the revised Framework, Section 3.3.1.3 stresses the need to for
consistency in the dose metric in the exposure assessment and the
effects assessment, which the peer reviewers raised as an important
issue (see Charge Question C.1.7).  Section 3.3.1.3 now also explains
that although tissue concentrations are the preferred dose metric, since
TEFs for mammals are largely derived from studies using administered
dose, application of mammalian TEFs to the diet is a more accurate
approach and will minimize variability in the analysis.  Therefore, the
mammalian diet example (Table 6) is prefaced with a discussion regarding
the fact that although tissue concentrations are the most relevant
dose-metric, it is often impractical or impossible to define dose on a
tissue-specific basis for mammals.  

Additional discussion of the limitations and variability associated with
calculating TECs based on diet rather than tissue concentrations (i.e.
internal dose) has been added to Sections 3.2.1.1 and 3.3.1.3 (see also
response to charge question C.1.7).  

The use of bioaccumulation factors to estimate tissue concentrations
from environmental media (or to relate known tissue concentrations back
to ambient levels) is described in section 3.3.1.4.  This section is
clearly written until the p. 35-p. 40 transition, at which it appears
that some words are missing.  In addition, the description of sediment
water concentration quotients (IIsocw ) and Di/r on pp. 40-41is somewhat
cryptic.

EPA Response: The Characterization of Exposure section of the Framework
(Section 3.3.1) has been reorganized and revised extensively to improve
clarity.  In addition, definitions of IIsocw and Di/r have been added to
Text Box 5.  

 

It is recognized that the WHO factors are starting points. From a
management perspective, it would be useful to have more discussion about
what situations “trigger” an assessment to develop
assessment-specific RPF values. The text should be enhanced to show how
to make these site-specific selections without being arbitrary and
without simply adopting the selections that are easiest, favored by the
entity that complains the most in the situation, or happen to be on the
computer at the time of the calculation.  Again, EPA needs to provide
more text with guidance on how to make this decision to reduce the
potential for arbitrary outcomes.

EPA Response:  Whether to use the TEF methodology and whether to use
consensus WHO-TEFs or RPFs are decisions specific to individual ERAs and
hence need to be determined on a case-by-case basis during planning and
problem formulation.  Sections 3.1 and 3.2 discuss issues for risk
assessors and risk managers to consider in the decision-making process. 
Section 3.3 and especially Section 3.3.2 (and the examples in Section
3.3.2.4), provide guidance on how to logically organize available ReP
data to make site-specific selections that minimize uncertainties and
maximize species similarity, endpoint and dose relevance, and
consistency.  Using this approach facilitates the transparent and
defensible selection of relative potency factor(s) for use in risk
assessments.

A suggestion was made that EPA consider the Bursian et al. (2003) paper
along with the Tillitt paper for the example on mink. Giesy provides a
rationale for this. 

EPA Response:  EPA believes the use of the Tillitt et al. (1996) paper
in the mink example is adequate for this Framework, as the example is
only intended to illustrate the procedure for organizing data and
selecting relative potency values and is not intended to evaluate the
toxicity or the TEF for mammals in general or for mink specifically.

A few of the reviewers found the examples for birds and for mammals
unclear. It may be helpful to have these read over by someone unfamiliar
with the methodology in order to identify how these examples can be made
more understandable.

EPA Response:  Section 3 in general, and the mink example in particular,
has been revised to improve clarity. 

It would be helpful to include a website address in the Framework
Document for the 1997 TEF database.  This database consists of all
relevant toxicological data for dioxin-like compounds through 1997.  It
was used to establish the WHO98 TEFs for fish, birds, and mammals given
in Table 2. It seems like there is more data available on RePs for
different species of birds based on embryo toxicity than is referenced
in the Framework Document.  It would be helpful to update the bird RePs
accordingly.

EPA Response: The 1997 ReP database created by the Karolinska Institute
is available on EPA’s web site along with the final Framework. To
EPA’s knowledge, the Karolinska Institute is not presently updating or
maintaining this database. However, the database of mammalian RePs was
reviewed and refined by Haws et al. (2006). This database was used in
the 2005 WHO reevaluation of mammalian TEFs as described in van den Berg
et al. (2006). The Haws et al. database is published in peer-reviewed
literature (Haws et al., 2006).  

 

Some simple ways to clarify the discussion of the matrix include: a)
just refer to it as the matrix (not the matrix model), b) refer to all
categories as “levels” and not “tiers” in order to distinguish
between these levels of information and tiers of risk assessment, c) P.
49, Lines 10 through 16. Simplify all of this by simply introducing the
Matrix as a tool for guiding the selection of ReP values from which to
derive a RPF.

EPA Response:  The suggested changes have been made in the revised
Framework.  

The dose specificity axis of Figure 10 is an important part of the
matrix.  However, this axis actually combines two different components
related to the dose metric (or exposure metric) used to determine RPFs. 
This is noted in the draft document [p. 52 lines 22-25] but the
discussion of these two aspects could be clarified and additional
guidance provided on how to balance these two components in the
selection of RPFs.  The first component is the degree to which the dose
metric used to derive RPs is the same as the dose metrics used in the
exposure assessment and in the effects assessment.  The authors call
this “consistency”.  The second component of this axis is the degree
to which the dose metric used to derive RPs is relevant to the target
tissue and effects of concern.  It is this component that is actually
reflected in the “tiers”: dose in tissue, dose in organism,
administered dose, and nominal/predicted dose.  The authors call this
“specificity”; “relevance” may be a better term.  In the
presentation of example 3 (mink; pp. 55-58) the authors point out a
situation in which a less relevant dose metric (administered dose) may
be preferable when it is more consistent with the dose metric used for
the effects assessment (TCDD dose-response curve).  The authors could
make a more explicit statement to provide additional guidance on how to
balance these two considerations.  For example, they might say that one
should choose RPs generated using the most relevant dose metric that is
also fully consistent with the dose metric used for the effects
assessment (i.e. consistency is given priority over relevance).

EPA Response:  The terminology used to label the z-axis of the Matrix
has been changed in the revised Framework to “Dose Relevance and
Consistency.”  Further, discussion of the z-axis has been re-written
to clarify the two aspects of dose under consideration when selecting
relative potency values to be used in deriving RPFs (see Section 3.3.2.3
entitled RPF Dose Relevance for Effect and Consistency with
Dose-Response Relationship).  To further this point, the Matrix and
discussion now include a strategy for weighing each of the components by
introducing a scale on the z-axis that facilitates summing of ReP dose
relevance with dose-response consistency.  

During the December 5th conference call there was a discussion of how
the Matrix could be made more clear. During that call, Peterson
recommended that the Matrix in Figure 10 be changed as highlighted
below: 

For the Y Axis, Endpoint Similarity the levels would be named: 

Toxic Effect of Concern in vivo

Other Toxic Effect in vivo

AhR-Dependent Biochemical Endpoint in vivo

AhR-Dependent Biochemical Endpoint in vitro

Other Biochemical Endpoints (AhR Binding)

Quantitative Structure Activity Relationships (QSAR)

For the X Axis, Species Similarity, Level 3 would be Vertebrate
Class-Specific "Consensus" TEFs

The Z Axis would be identified as Target Tissue Similarity / Dose
Similarity

EPA Response: The y-axis levels have been revised, largely as suggested,
although they have been abbreviated in some cases in order to fit into
the figure.  The x-axis change was not incorporated because the use of
“Consensus TEFs” would imply this level would always be the WHO
TEFs.  However, EPA envisions that when additional data are available
and relevant, a risk assessor may select among any existing RePs to
derive RPFs and would not necessarily have to be constrained to
selection of the consensus TEFs.  The z-axis was revised as described
above in response to comment C.5(b).2 to clarify and highlight the two
aspects of dose similarity to be considered.

Following the conference call, Mark Hahn provided the following
additional commentary on the Matrix:

The y-axis might best be called “Endpoint relevance” (referring to
its relevance to effects of greatest concern).

The x-axis should be called “Species similarity” as suggested by
Dick.

The z-axis should be called “Dose metric consistency and relevance”
to reflect the two aspects of this axis, as discussed above.

EPA Response:  Changes consistent with these suggestions have been made
in the revised Framework.

On P. 59, Line 32, a key point is made that needs to come earlier in the
section and certainly at the beginning of 3.3.2.4. That point is that
you start with the TEFs and only become more site or species specific
when there is very good reason. Further, as more information becomes
available, the Matrix can be used to guide the development of new
default TEF values.

EPA Response:  Text expressing the expectation that WHO-TEFs can and
will be used in ecological risk assessment has been added to the
beginning of Section 3.3.2.  

 

The influence of detection levels on the uncertainty around risk
estimates could be addressed in the uncertainty section. 

EPA Response:  The suggested addition has been made in the revised
Framework.

It would be helpful to have a bit more information on the relative
magnitudes and direction of uncertainties around estimates. It may be
helpful to have discussion around the uncertainties associated with
selection of BSAFs (or other methods for estimating bioaccumulation)
relative to the uncertainties around TEF and RPF values. It may be
helpful to encourage users of this document to use sensitivity analyses
to guide the levels of effort they devote to the different components of
applying the TEQ/TEF methodology. Not all aspects of the methodology
have similar degrees of variability and uncertainty nor do they have an
equivalent impact on the final outcome the TEF methodology.

EPA Response: The uncertainty associated with a risk estimate will be
specific for each individual risk assessment and dependent on each
component (e.g. chemical concentration, BAF or BSAF, RPF or TEF, TRV) of
the risk estimate.  Therefore it is not possible to provide a relative
measure of uncertainty for BSAFs vs. TEFs or RPFs.  The selection matrix
approach described in Section 3.3.2 is designed to guide risk assessors
through evaluation of available relative potency data to, in part,
identify where the greatest uncertainties in relative potency data lie. 
In addition, Section 3.4.3 has been re-organized in an attempt to prompt
risk assessors to think about uncertainties associated with the use of
the TEF methodology itself within the broader context of an ecological
risk assessment (i.e. that the methodology serves as a framework for
conducting a qualitative sensitivity analysis).  

Section 3.4.3.1.1 suggests that there are non-AhR-dependent mechanisms
of action, but is vague on the point.  There are certainly
non-AhR-dependent mechanisms known in the toxicology literature, and the
section must point that fact out, give at least some mention of which
ones (immune systems, neurological, developmental, estrogenic) are known
and offer something more in the way of explanation.  This uncertainty
would underestimate the effects of these compounds. Section 3.4.3.1.2
refers to no known interactions, yet Cook et al. in Rolland et al., 1998
report synergistic responses in fish from exposure to TCDD and PCBs. 
Section 3.4.3.1.4 refers to the TEFs and RPFs as point estimates, yet
fails to acknowledge that these point estimates were the result of a
consensus meeting among scientists form different countries. Point
estimates work with little uncertainty if there is a huge database to
support them (and a low C.I.) or if they are set as protective, as in a
barrier.  However, these point estimates are neither.  There is but a
modest database and no attempt to set these as “not greater than” in
regulatory terms.  Therefore, one source of error/uncertainty is the
greater response (or lesser) due to the biological differences among
animals for the same species, or genus or family or even order.  These
basic biological differences could account for huge uncertainty and
natural variation.

EPA Response:  Language has been added to the Preface and the
Introduction to clarify that the Framework is not intended to provide
comprehensive guidance on conducting ecological risk assessment.  The
purpose of the Framework is to provide guidance on how to apply the
toxicity equivalence factors for dioxin-like activity of chemicals
within an ecological risk assessment for dioxin-like chemicals. 
Accordingly, the Framework does not include guidance on determining or
accounting for ecological affects from other modes or mechanisms of
action, whether from PCDDs, PCDFs, and PCBs or other chemicals.  

Not accounting for all possible modes or mechanisms of action of PCDDs,
PCDFs, and PCBs (or any other chemicals) will not necessarily result in
an underestimate of effects or risks.  The relative potency of the
various modes/mechanisms of action relative to the exposure
concentrations would need to be considered.  For dioxin-like PCDDs,
PCDFs, and PCBs, current evidence indicates that the greatest potential
for effects on ecological endpoints of most concern (e.g., growth,
survival, reproduction) is from the AHR agonists (Giesy and Kannan,
1998; Rice et al., 2002).  Nonetheless, determining the chemicals,
modes/mechanisms of action, species, and endpoints of concern is
assessment-specific and is therefore best performed during the planning
and problem formulation phases of the individual risk assessment.

The methods used to estimate tissue levels are likely to have the
greatest uncertainties associated with them. Because there are various
methods by which tissue residues can be measured or estimated, the
Framework should expand on this source of uncertainty in the application
of the method. This is discussed further under Charge Question 8. Giesy,
Metcalf, Kennedy, Hope and Menzie provide detailed discussion on this
issue.

EPA Response:  A discussion of uncertainties associated with the
characterization of exposure, including both measuring and estimating
tissue concentrations, is included in Section 3.4.3.2.2 of the revised
Framework. In addition, additional discussion and references regarding
the use and/or extrapolation of BAFs/BSAFs and food-web modeling have
been added to Section 3.1.5.

One issue not addressed specifically concerns some of the uncertainties
and complexities associated with the additivity assumption.  For
example, the issue of ligand “intrinsic efficacy” and how it
(together with ligand affinity) contributes to the “potency” of AHR
agonists is not mentioned.  The issue may be too technical to treat in
this Framework (e.g. on p. 10), but it is relevant to the additivity
assumption in that compounds with lower intrinsic efficacy can act as
“partial agonists” and thus inhibit the response to full agonists at
certain dose ratios (Toxicol. Appl. Pharmacol. 168: 160).  This has been
shown both theoretically and experimentally, but the extent to which it
occurs with environmentally relevant mixtures is not clear.

EPA Response:  Additional text summarizing empirical data from both
laboratory and field studies on ecological species that provide strong
support for the additivity model has been added to Section 2.1.  This
section also includes reference to U.S. EPA, 2000a, as it has an
extensive discussion regarding the empirical research and
“receptor-based theory” underyling the additivity assumption.  This
information was not reiterated in the Framework in the interest of
keeping the document clear and concise and focused on the application of
the TEF Methodology.  Discussion regarding “intrinsic efficacy” and
the potential for partial agonists to act as antagonists at
environmental concentrations has also been added to Section 2.1 and
3.4.3.1.2 by noting that   SEQ CHAPTER \h \r 1 Van den Berg et al.
(1998; 2006) concluded that antagonistic effects are usually seen above
environmentally relevant doses, such that the presence of chemicals that
have demonstrated antagonist activity (primarily in vitro) is unlikely
to result in large errors when antagonists are present.  

The uncertainty section should include some discussion regarding the
source information for derivation of RePs. RePs determined from NOAELs,
LOAELs, and benchmark doses are not as accurate as those based on LC50s,
EC50s, LD50s or ED50s. 

EPA Response: Language consistent with this suggestion has been added to
Section 3.4.3.1.3 in the revised Framework.

Bioanalytical tools are identified on P. 66, Line 16 as a means of
reducing uncertainty. But earlier these tools were referred to as
screening tools and not ready for risk assessment. This may need further
discussion with regard to how and when these tools can be used to
address uncertainty.

EPA Response: The Framework notes that the experts at the EPA/DOI
workshop concluded that such bioanalytical tools should not be used as
an alternative to congener-specific analysis and the toxicity
equivalence methodology.  Rather these bioanalytical analyses are
complementary tools that can be useful in providing additional lines of
evidence. The referenced sentence has been revised to clarify that
bioanalytical tools may reduce uncertainty by providing another line of
evidence regarding whether dioxin-like toxicity risks are fully
represented by the TEFs-WHO98/05.

 

Reviewers typically provided suggestions for references within the
context of specific comments. EPA should review these for contextual
information. 

EPA Response:  Many of the suggested references have been incorporated
in the revised Framework.  In some cases, a more recent or more relevant
reference on the topic was included rather than the one suggested.

 

The types of methods by which exposures (in the diet or in the tissues)
can be measured or estimated. The Framework restricts itself largely to
discussing this in terms of “factors” such as BAFs and BSAFs. Such
factors are one of several ways by which exposure information can be
developed. The other two important means are direct measurement and the
use of bioaccumulation and food-chain models. These might include steady
state as well as kinetic models. During our conference call, it appeared
that BSAF was being used to imply the use of all of these tools.
However, this will lead to confusion on the part of practitioners who
think of BSAFs as factors (e.g., taken from a table or derived to
reflect steady state conditions). The use of measurements and models do
not receive adequate discussion in the framework.  The discussion of
exposure within the Framework can easily be broadened to be inclusive of
the various methods available for estimating exposures and doses and not
to indicate that the method is exclusively related to selection of BSAF
or BAF factors. See Menzie for suggestions on where changes can be
easily made to accommodate this larger view. Also, during our conference
call, Phil Cook indicated that there was some information that could be
added to help the reader work through the proper selection of methods
and/or to have confidence in certain values. 

EPA Response:   See response to comment C.4.1. 

Many comments were made concerning the application of BSAFs and BAFs.
These fall into several categories. Collectively the comments suggest
that this part of the TEF/TEQ approach can use some careful re-working.
This might be reduced as an issue if BSAFs are subsumed into a broader
discussion of measuring and/or estimating body burdens. BSAFs then are
but one tool that can be used and not the only tool. 

EPA Response:  See response to comment C.4.1.

There is no suggestion of a reliable, non-controversial source of
universally applicable “generic” BSAF values which would allow this
approach to be used in lieu of site-specific information.  Much more
needs to be said about where or how one obtains the BAFs/BSAFs essential
to the application of this method.  It also needs to be made clear
whether the BSAF values in Tables 4-6 are intended as examples only or
as de facto “generic” factors.  The challenges associated with
measuring BAFs/BSAFs are also understated here.  The Group generally
felt that “extrapolation” is a non-controversial way around any of
these challenges.

EPA Response:  The Framework does not suggest a source of universally
applicable “generic” BSAF values because EPA does not advocate or
support such an approach.  EPA’s approach for acquiring and using
BAFs/BSAFs is summarized in Section 3.3.1 of the Framework and is based
on many previously published peer-reviewed publications (Burkhard, 2003;
Burkhard et al., 2003a, b; 2004; 2006) and EPA guidance documents (U.S.
EPA, 1993; 1995a, b, c; 2000; 2001a; 2003b).  Indeed, EPA’s
bioaccumulation approach includes extrapolation of BAFs/BSAFs when
appropriate conditions are met and appropriate normalizing factors are
incorporated (U.S. EPA, 1993; 1995a, b, c; 2000; 2001a; 2003b).  This
body of information is summarized rather than reiterated in detail in
the interest of keeping the Framework concise and focused on the
application of the TEF Methodology.  However, additional references to
previously published peer-reviewed articles, EPA guidance, and EPA’s
BSAF data set have been added to Sections 3.3.1.4 and 3.3.1.5.  A note
has also been added to Tables 4 – 6 to clarify they are not intended
to be “default” values, and a note has been added to each table to
clarify this point. 

If the use of BSAFs is to be advocated, there should be more discussion
of the assumptions of the technique and the range of expected values and
the limitations of the technique.

EPA Response:  EPA’s approach for acquiring and using BAFs/BSAFs, as
summarized in Section 3.3.1 of the Framework, is well established and
based on many previously published peer-reviewed publications (Burkhard,
2003; Burkhard et al., 2003a, b; 2004; 2006) and EPA guidance documents
(U.S. EPA, 1993; 1995a, b, c; 2000; 2001a; 2003b); all of which are
referenced in Section 3.3.1.  This body of information is extensive.  In
the interest of keeping the Framework concise and focused on the
application of the TEF Methodology, the approach is summarized rather
than reiterated in detail; however, additional references to these
articles and guidance documents have been added Sections 3.3.1.4 and
3.3.1.5 in the revised Framework. 

One reviewer suggested that the statements on the limitations of the use
of TEC in the diet be made more apparent.  While the discussion points
out these limitations, it comes to the conclusion that this is an
acceptable practice when additional information is not available.  It is
this reviewer’s opinion that the concentrations in target tissues
should be predicted with congener-specific BAF or BMF values and then
the TEFs applied to calculate predicted tissue-specific TEC
concentrations which can then be compared to toxicant reference values
(TRVs). Because of associated uncertainty, it would be useful to
highlight the value of multiple lines-of-evidence approaches.

EPA Response:  EPA believes that it is appropriate, in some cases, to
use diet-based TECs. The rationale and limitations for this approach for
mammals is discussed in Sections 3.3.1.3, 3.3.1.5, and 3.3.2.4.3. 
Example 3 also discusses how multiple lines-of-evidence (tissue-based
TRVs and diet-based TRVs) could be compared, when appropriate data are
available.    

There was a strong sentiment among reviewers that BAFs (water to tissue)
would not be a reliable way to estimate tissue levels. For example, the
report includes an admission that dioxins, furans and non-ortho PCBs
would be present in water under most exposure scenarios at
concentrations well below detection limits. Data are rarely available on
the ng/L concentrations of these hydrophobic compounds in water, since
this would require extraction of large volumes of water. While part of
the concern relates to the ability to estimate or measure the
concentrations of dioxin-like compounds in water, there is also a
concern that empirical BAF values may be highly variable and contribute
to substantial uncertainty in exposure estimates. Metcalfe and Giesy
give detailed discussion of these concerns.

EPA Response:  While it has historically been difficult to measure low
concentrations of PCDDs, PCDFs, and PCBs in water, as acknowledged in
the Framework, it is possible and increasingly feasible to perform such
measurements given newer analytical techniques.  In fact, high quality
BSAFs have been measured in a number of ecosystems (U.S. EPA 1995a;
Burkhard et al., 2004; see also EPA’s BSAF data set at
http://www.epa.gov/med/Prods_Pubs/bsaf.htm).  As summarized in Section
3.3.1.4, EPA has developed extensive guidance for minimizing variability
in BAF and BSAF measurements and when extrapolating BAFs and BSAFs
across ecosystems with similar conditions (U.S. EPA 1995a; 2000; 2003b).
 Also included in Section 3.3.1.5 is a discussion of approaches (e.g.
the use of food-chain models and/or the “hybrid modeling approach”)
that can be taken to adjust BAFs/BSAFs to decrease variability and
increase accuracy when extrapolating across ecosystems.

With regard to BSAFs, there are several technical issues related to the
application of BSAFs for predicting tissue concentrations that were not
discussed in sufficient detail in the Framework.

The concentrations of chlorinated contaminants in sediments are
typically very heterogeneous; both vertically with sediment depth and
horizontally in river or lake ecosystems. The sediment concentration
chosen for the risk analysis exercise will be critical to the outcome,
but no guidance is provided on the solution to this challenge.

EPA Response:  The Framework is not intended to be a comprehensive guide
for conducting ecological risk assessment for dioxin-like chemicals. 
Sampling designs and collection methods are not issues specific to the
use of the TEF methodology, but rather issues to be addressed in the
analysis plan for a risk assessment of dioxin-like chemicals. 
Nonetheless, EPA’s extensive guidance for minimizing variability in
BAF and BSAF measurements (U.S. EPA 1995a; 2000; 2003b) is referenced in
the section that summarizes EPA’s approach for measuring and
extrapolating bioaccumulation factors for PCDDs, PCDFs, and PCBs
(Section 3.3.1.4).

The Framework currently suggests that BSAFs can be used to predict the
concentrations of chlorinated contaminants in fish from concentrations
in sediment.  An example is provided using BSAF data for Lake Ontario. 
There may be enough data in the literature from various aquatic
ecosystems to generate reasonable estimates of the sediment/fish BSAFs
for the many of the dioxin, furan, and PCB congeners (although this is
subject to debate). However, there are few data in the literature on
BSAFs calculated from the ratio of contaminant concentrations in
sediments and the eggs of fish-eating birds. The report provides BSAFs
calculated from sediment and herring gull egg data for the Lake Ontario
ecosystem, but applying these BSAFs to other ecosystems (e.g. rivers,
shallow lakes, etc.) would/could introduce substantial uncertainty. With
respect to this potential uncertainty, the report should identify other
approaches for determining the residues of chlorinated contaminants
including direct analysis of bird eggs.

EPA Response:  The Framework highlights the strengths of measuring
tissue concentrations (residues or internal dose) of PCDDs, PCDFs, and
PCBs in Sections 3.2.1.1 and 3.2.1.2.  Section 3.3.1.3 highlights the
accuracy of using TECs based on measurements of PCDDs, PCDFs, and PCBs
in tissues.  Section 3.3.1.4 begins with a discussion highlighting the
fact that the most straightforward way of calculating TECs is from
measured concentrations of dioxin-like chemicals in tissues using
Equation 2-1.

The limitations of applying a BSAF to estimate tissue residues have not
been adequately described.  The Framework does not address the
variability and precision inherent in this approach relative to
predictions of contaminant concentrations in flora and fauna within
ecological systems or between ecological systems.  Thus, the magnitude
of potential errors generated in predicting contaminant concentrations
in wildlife and plants can not be put into perspective relative to other
sources of variability and uncertainty that are inherent in the TEF
methodology.  In part, this is due to the reliance of these models on
lipophilicity as the only determinant of accumulation.  However, studies
have shown that this factor alone is not a sufficient predictor of
bioaccumulation and in fact, accumulation is a function of many factors
including molecular size, conformation, sediment characteristics and
biological factors (feeding habits).

EPA Response:  EPA’s approach for acquiring and using BAFs/BSAFs, as
summarized in Section 3.3.1 of the Framework, are well established and
based on many previously published peer-reviewed publications (Burkhard,
2003; Burkhard et al., 2003a, b; 2004; 2006) and EPA guidance documents
(U.S. EPA, 1993; 1995a, b, c; 2000; 2001a; 2003b).  This body of
information establishes that bioaccumulation of PCDDs, PCDFs, and PCBs
is a function of many factors (e.g. trophic level, food web
characteristics, sediment organic carbon, organismal lipid, and
sediment-water concentration quotient) and provides in-depth discussion
of the uncertainties associated with the use of BAFs/BSAFs for
estimating tissue concentrations.  However, this body of information is
extensive and in the interest of keeping the Framework concise and
focused on the application of the TEF Methodology, the uncertainties are
summarized, rather than reiterated in detail, in Section 3.4.3.2.2 in
the revised Framework. 

If BMFs and BSAFs are used to predict concentrations of PCDD/DF in
tissues, an upper and lower bound could/should be given for the
concentrations of each congener and this range of values propagated
through the calculation of the TECs in tissues.  To this end,
probability bounds may be a useful tool.

EPA Response:  EPA agrees that providing a range of values (upper and
lower bound) for the congener concentrations would be useful to
illustrate the variability around these values.  However, in the
interest of keeping the example TEC calculations concise and simple for
the purpose of illustration and to avoid the misperception that the
values are anything other than hypothetical (a concern expressed by the
reviewers regarding the values in Tables 4 – 6; see comment #B.2), a
range of values have not been incorporated into Tables 4, 5, or 6.

One reviewer suggested including an approach (either a description of
method or an example) that would serve to illustrate how the TEF/TEQ
approach could be validated. He notes that while the examples are
illustrative, he would prefer to see a kind of validation for this
approach with real ecological situations indicating the feasibility and
possible uncertainty. He suggests two exercises: 

Model the transfer of dioxin-like compounds from actual sediment
concentrations with the endpoint being a prediction of concentrations
for species higher in the food chain. These data could than be compared
with actual concentrations found in the relevant species for that
specific environmental situation. 

The second validation could be done in a reverse way. In this case
calculations should go back from TEC concentrations observed in an
actual top predator species and calculate the possible concentration
levels in species at lower trophic levels and the sediment. 

Both exercises should produce more clarity about the predictive power of
the suggested EPA method described in chapter 3.3.1.4. 

EPA Response:  EPA has performed the first exercise (Cook et al., 2003).
 Reference to this peer-reviewed publication has been added in
appropriate sections (Sections 3.3.1.3 – 3.3.1.5) of the revised
Framework to demonstrate/clarify the predictive power of the TEF
Methodology.  Numerous studies have demonstrated the second type of
“validation,” i.e. correlations between effects of environmental
mixtures in marine mammals and avian species and food-web dietary
concentrations (Ross et al., 1996; Summer et al., 1996a, b; Giesy and
Kannan, 1998; Restum et al., 1998; Shipp et al., 1998a, b; Ross, 2000). 
References to these studies have been added to Section 2.1 of the
revised Framework. 

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Burkhard, LP; (2003) Factors influencing the design of BAF and BSAF
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Burkhard, LP; Cook, PM; Mount, DR. (2003a) The relationship of
bioaccumulative chemicals in water and sediment to residues in fish: a
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Burkhard, LP; Endicott, DD; Cook, PM; Sappington, KG, Winchester, EL.
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Burkhard, LP; Cook, PM; Lukasewycz, MT. (2004) Biota-sediment
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Burkhard, LP; Cook, PM; Lukasewycz, MT. (2006) A hybrid
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Burkhard, LP; Cook, PM; Lukasewycz. MT. (2008) Organic Carbon -Water
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  PAGE  2 

Charge Question 1: A main goal of this document is to assist ecological
risk assessors in applying the toxicity equivalence methodology
correctly. Please comment on the overall effectiveness of the document
in achieving this goal. Please discuss document organization,
appropriateness of the level of detail, and usefulness of
figures/tables.

Charge Question 2: The document proposes to resolve current
inconsistencies in the scientific literature over terms such as
“ReP” by establishing and using clearly-defined, unified terms.
Please comment on the clarity and effectiveness of the terms used.

Charge Question 3: Please comment on whether the advantages of using the
toxicity equivalence methodology are adequately explained.

Charge Question 4: The framework emphasizes the importance of measuring
or estimating chemical-specific PCDD, PCDF, and PCB concentrations in
tissues in order to apply the methodology. Please comment on this and
whether sufficient discussion of estimating concentrations in tissues is
provided. Is the explanation of the application to the methodology to
dietary exposure in mammals, as distinguished from fish and birds,
adequate?

 Charge Question 8: Is the discussion of exposure and bioaccumulation
sufficient for basic applications of TEFs and RPFs in ecological risk
assessments? Please explain.

Charge Question 7: Are you aware of any essential references that have
been omitted?

Charge Question 6: Please comment on whether the uncertainties
associated with the application of the toxicity equivalence methodology
are comprehensive and adequately explained.

Charge Question 5(b): The framework provides considerations for
selection of relative potency factors that may be more specific for the
species, endpoints, and doses of concern in individual ecological risk
assessments than the international consensus TEFs. Are the matrix
presented in Figure 10 and the examples used to illustrate the
application of the matrix clear and adequately explained? Are there
elements which should be added or removed from the matrix? Do you agree
with their place in the tiers on the matrix? Please explain.

 

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assessments than the international consensus TEFs. Please comment on the
completeness and clarity of this discussion.