INTERIM CONSENSUS
OF THE HARS SCIERTIFIC
3rwuticaJ
mile limit
Phase 1: Human Health Evaluation
Proposed Bioaccumulation Testing Evaluation Framework for
Assessing the Suitability of Dredged Material to be Placed at the
Historic Area Remediation Site (HARS)
Contract No. 68-C-00-121
Work Assignment No. 2-35
Prepared for
U.S. Environmental Protection Agency
Region 2
New York, NY
June 20, 2002
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Duxbury Operations
397 Washington Street
Duxbury, Massachusetts 02332
Telephone 781-934-0571
Fax: 781-934-2124
June 20,2002
Mr. Ronald Borsellino
USEPA, Region 2
290 Broadway
New York, New York 10007-1866
Mr. Bryce Wisemiller
USAGE, NY District
Jacob K. Javits Building
26 Federal Plaza
New York, NY 10278-0090
Dear Mr. Borsellino and Mr. Wisemiller:
On behalf of the HARS Scientific Peer Review Panel, Battelle is pleased to submit the enclosed document
entitled "Interim Consensus Report of the HARS Scientific Peer Review". This document represents the
consensus responses prepared by the HARS Peer Review Panel (Panel) to the charges presented on January 10
and 11,2002 for Phase I of the HARS Scientific Peer Review. This phase of the Peer Review focused on issues
associated with the development of bioaccumulation threshold values protective of human health.
Please feel free to call me at 781-952-5384 if you have any questions or would like clarification from the Panel
regarding specific responses.
Sincerely,
Nancy L. Bonnevie
Enc/
Cc: Dr. Carlton Hunt, Battelle
RMW Members
Panel:
Dr. Allen Burton
Dr. Ken Jenkins
Dr. Peter Landrum
Dr. James Meador
Dr. Lynn McCarty
Dr. Anne McElroy
Dr. Harlee Strauss
Mr. Paul Price
Mr. Richard Wenning
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INTERIM CONSENSUS REPORT
Phase 1: Human Health Evaluation
Proposed Bioaccumulation Testing Evaluation Framework for
Assessing the Suitability of Dredged Material to be
Placed at the Historic Area Remediation Site (HARS)
Contract No. 68-C-00-121
Work Assignment No. 2-35
Prepared for
U.S. Environmental Protection Agency
Region 2
New York, NY
Prepared by
Battelle
397 Washington Street
Duxbury, MA 02332
(781) 934-0571
June 20,2002
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HARS Peer Review: Responses June 20,2002
Page i
TABLE OF CONTENTS
EXECUTIVE SUMMARY iii
1.0 INTRODUCTION 1
1.1 Selection of Panel Members 1
1.2 Peer Review Process 2
1.3 Phase I Document Organization 2
2.0 RESPONSE TO CHARGES 3
2.1 Response to Charge 1 3
2.2 Response to Charge 2 7
2.3 Response to Charge 3 8
2.4 Response to Charge 4 9
2.5 Response to Charge 5 9
2.6 Response to Charge 6 9
2.7 Response to Charge 7 10
2.8 Response to Charge 8 11
2.9 Response to Charge 9 11
2.10 Response to Charge 10 12
2.11 Response to Charge 11 13
2.12 Response to Charge 12 14
2.13 Response to Charge 13 15
2.14 Response to Charge 14 15
2.15 Response to Charge 15 15
2.16 Response to Charge 16 16
2.17 Response to Charge 17 16
2.18 Response to Charge 18 17
2.19 Response to Charge 19 20
2.20 Response to Charge 20 20
2.21 Response to Charge 21 21
2.22 Response to Charge 22 22
2.23 Response to Charge 23 23
2.24 Response to Charge 24 23
2.25 Response to Charge 25 23
2.26 Response to Charge 26 23
2.27 Response to Charge 27 24
2.28 Response to Charge 28 24
2.29 Response to Charge 29 24
2.30 Response to Charge 30 26
3.0 REFERENCES 27
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LIST OF ACRONYMS
Ag
silver
Cd
cadmium
CRM
certified reference material
Cu
copper
DQO
data quality objective
GC
gas chromatography
HARS
Historic Area Remediation Site
HEP
Harbor Estuary Program
IEUBK
integrated exposure uptake/biokinetic model
log Kow
octanol water partitioning coefficient
MDL
method detection limit
MDS
Mud Dump Site
MLE
maximum likelihood estimate
Ni
nickel
NJMSC
New Jersey Marine Sciences Consortium
NOAA
National Oceanic and Atmospheric Administration
NS&T
National Status and Trends Program
NY/NJ
New York/New Jersey
OAPCA
Organotin Antifouling Paint Control Act
PAHs
polycyclic aromatic hydrocarbons
Pb
lead
PCBs
polychlorinated biphenyls
QAJQC
quality assurance/quality control
QSAR
Quantitative Structure Activity Relationship
RfDs
reference dose
RMW
Remediation Materials Workgroup
ROS
regression order statistics
TBT
tributyltin
TCDD
tetrachlorodibenzo-p-dioxin
TDI
total daily intake
TEF
toxicity equivalent factors
TEQ
toxic equivalence
USACE
United States Army Corps of Engineers
USEPA
United States Environmental Protection Agency
Zn
zinc
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EXECUTIVE SUMMARY
Under the direction of Battelle, a Scientific Peer Review Panel (Panel) was charged on January 8-9,2002
with review of the December 21, 2001 draft report prepared by the United States Environmental
Protection Agency (USEPA) Region 2 titled, "Proposed Bioaccumulation Testing Evaluation Framework
for Assessing the Suitability of Dredged Material to be Placed at the Historic Area Remediation Site
(HARS)" (hereafter referred to as the HARS Framework). The Panel was asked to provide technical
comments on thirty questions and concerns developed by USEPA Region 2 (EPA) and the Remediation
Materials Workgroup (RMW) pertaining to evaluation of the potential human health risks associated with
evaluation of the suitability of dredged sediment for use as remediation material at the HARS. Questions
and concerns pertaining to ecological issues will be addressed in a second phase of the peer-review
process. This document addresses questions pertaining to human health considerations.
Technical Findines
Overall, EPA is congratulated by the Panel for undertaking a process to revise the current interim
program for evaluating the suitability of dredged material for use as remediation material at the HARS. It
is evident that EPA strove to develop a scientifically defensible evaluation process. While EPA
succeeded in several aspects, the Panel concluded that the overall framework requires considerable
revision with regard to the calculation of HARS-Specific values protective of human health. Additional
refinements of the proposed HARS Framework are needed to ensure that the most current understanding
of fate and effects processes and the most advanced practical procedures for evaluating contaminant
migration through the food web and human health risks are included in the final framework.
The Panel identified several general concerns, which EPA is urged to consider and address in the final
dredged material evaluation framework:
1. There is a need for greater clarity and transparency in the evaluation framework. EPA should
clearly describe the logic and scientific evidence used in the development and parameterization of
the models used. In addition, the Panel agreed that focusing the model on uptake for a few key
species associated with the HARS will enhance the clarity and transparency while reducing the
uncertainties of the evaluation.
2. The receptor(s) to be protected in the conceptual model used in the human health risk assessment
must be carefully defined. As currently described, it is unclear whether the assessment is based on
a typical fish consumer or a high-end fish consumer. In addition, only chronic health effects are
evaluated; for fish ingestion the effects of short-term exposure during sensitive periods (e.g., fetal
development) may be the critical consideration. Finally, the issue of potential future use is not
addressed.
3. There is a need to improve the characterization of current environmental conditions at the HARS.
Currently there is insufficient and/or potentially outdated evidence concerning the status of the
fishery, recreational angler activity and catch success, contaminant residues in benthic organisms
and fish tissues, geochemistry, and the fate and effects of contaminants placed at the HARS.
Given the considerable resources invested in future remediation efforts, as well as the opportunity
to potentially improve conditions at the HARS while at the same time supporting port/harbor
development and the region's economy, it seems prudent that EPA implement a HARS-specific
sediment and biological characterization study and monitoring program to establish current,
baseline environmental conditions and to continually monitor for future changes. At present,
there appears to be little credible evidence upon which to ascertain whether (and how) the
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placement of dredged sediments will improve conditions (e.g., reduce contaminant body burdens
in fish) at the HARS.
4. The Panel recommended that more thorough and formal uncertainty and sensitivity analyses be
conducted. In addition, while stopping short of recommending a complete change from the
proposed deterministic approach, the Panel did suggest that the potential utility of probabilistic
techniques, such as Monte Carlo analysis, should be carefully considered when selecting specific
parameters. EPA is strongly encouraged to consult available guidance documents and to use a
probabilistic approach as a first step to addressing many of the technical deficiencies identified in
this review.
Going forward, the Panel encourages EPA to further refine the suite of sediment assessment tools
indicated in the HARS Framework, particularly thosR associated with evaluation of chemistry conditions
and biological impacts associated witK^oject sediipgnts. Efforts should focus on the introduction of
HARS-specific environmental and exposureTSctors into the data compilation, risk assessment and
sediment evaluation processes. EPA should recognize that whatever final form the framework takes, the
process will likely need to be revised again in the future, as more information becomes available. It is
expected that advances in the scientific understanding of chemical fate and effects in the marine
environment will continue. EPA should be prepared to assess and adopt new information and evaluation
methods, as appropriate, in the future.
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1.0 INTRODUCTION
Battelle was contracted by EPA to coordinate a scientific peer review of the Proposed Bioaccumulation
Testing Evaluation Framework for Determining the Suitability of Dredged Material to be Placed at the
Historic Area Remediation Site (HARS). The HARS was designated by the United States Environmental
Protection Agency (USEPA) on September 29, 1997 simultaneous to the de-designation and terminated
use of the New York Bight Dredged Material Disposal Site [also known as the Mud Dump Site (MDS)].
Prior to the HARS designation, EPA committed to conduct a public and scientific peer review process of
its dredged material testing evaluation procedure (New York/New Jersey Harbor Estuary Program
[NY/NJ HEP], 1996). Pursuant to that commitment, the EPA prepared a charge and presented their
testing evaluation framework to a scientific peer review panel. Based on the comments received from
that initial peer review, EPA proposed changes to the testing evaluation framework, and sought to have
another peer review of the suggested modifications, conducted in accordance with guidance provided in
EPA's Peer Review Handbook (EPA, 2000a).
Battelle is managing all aspects of the peer review, including selection of the Panel, briefing the Panel and
organizing meetings. The peer review is being conducted in two phases, with the first phase focusing on
technical issues related specifically to the calculation of values protective of human health as well as
those issues shared by both the human health and ecological evaluations. The second phase will focus on
those issues specifically related to the calculation of values protective of ecological receptors.
1.1 Selection of Panel Members
The Panel convened for the peer review was selected by Battelle from a list of candidates recommended
by the Remediation Material Workgroup (RMW) members. However, final selection of the Panel was
conducted independently by Battelle without input from EPA or any other RMW member. Nine
individuals were selected that had demonstrated expertise in one or more of the following disciplines:
• toxicology;
• aquatic toxicology;
• analytical chemistry;
• geochemistry;
• sediment assessment;
• fisheries biology;
• ecology;
• human health risk assessment;
• ecological risk assessment; and
• metals chemistry.
In addition to demonstrated expertise in one of these disciplines, each candidate was asked to confirm the
following to screen out any perceived conflict of interest in comments or viewpoints expressed during the
peer review:
• The candidate must not have, or have had, substantial involvement in federal or state regulatory
activities for dredging in New York and New Jersey Harbor.
• The candidate must not be under contract to any environmental or shipping organization with a
perceived interest in the results of any project potentially evaluated by the subject Testing
Evaluation Framework.
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• The candidate must not have had any membership or inferred involvement in any organization
that possesses extreme points of view on environmental or shipping interests.
• The candidate must not be perceived to possesses any bias or extreme points of view towards
either environmental or shipping interests.
• The candidate must not possesses any extreme points of view related to the subject matter or
scientific material being reviewed.
Brief descriptions of each of the panel members are provided in Appendix A.
1.2 Peer Review Process
Figure 2-1 provides a summary of the overall process that is being followed for each phase of the peer
review (Appendix C). Briefly, at the beginning of each phase, the Panel will be convened for a two day
meeting in New York City for a briefing on the peer review process and the delivery of Charges specific
to that phase. Representatives of EPA Region 2, United States Army Corps of Engineers (USACE), and
the RMW will be present to provide background information and perspectives on the various technical
issues addressed by the charge. For Phase I, this meeting was held on January 10-11,2000 (Appendix B).
Following the initial briefing meeting, each reviewer will have 60 days to independently review the
information provided and answer the specific charges relating to human health and shared issues. At the
end of this period, each reviewer will submit their independent responses to the charges to the Battelle
Peer Review Leader. All submissions will be in electronic (preferably MS Word) and hard copy format.
It is required that the reviewers refrain from discussing the charges and their responses with any member
of the RMW (including EPA and USACE) during this independent review period. Specific technical
questions and requests for clarification that arise during this period will be directed to the Battelle Peer
Review Leader. The independent responses received from each peer reviewer for Phase I are provided in
Appendix D.
All peer reviewers will reconvene for a two-day issue resolution meeting to discuss the draft narrative
report which will be compiled by Battelle based on the comments received from the independent reviews.
During the meeting, a facilitated discussion will be held to discuss outstanding issues and to resolve areas
of conflicting opinion.
Based on the outcome of the issue resolution meeting, an interim consensus report for each phase will be
prepared by the Panel and submitted to EPA, USACE, and the RMW. Upon completion of Phase II,
these interim reports will be combined into a final draft response, and a briefing meeting will held with
the RMW to discuss.
1.3 Phase I Document Organization
This document represents the interim consensus report for Phase I of the HARS Peer Review. Section 2
provides the consensus statements prepared by the Panel in response to each charge presented by EPA
and the RMW. Where necessary, background information is provided by the Panel to explain or justify
the consensus reached.
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2.0 RESPONSE TO CHARGES
Under the direction of Battelle, a Scientific Peer Review Panel (Panel) was convened in New York City
on January 10-11, 2002 and charged with the review of the December 21, 2001 draft report prepared by
USEPA Region 2 titled, "Proposed Bioaccumulation Testing Evaluation Framework for Assessing the
Suitability of Dredged Material to be Placed at the Historic Area Remediation Site (HARS)" (hereafter
referred to as the HARS Framework). This document focused specifically on issues pertaining to the
development of HARS-Specific Values for the protection of human health. The Panel was asked to
provide technical comments on thirty questions and concerns developed by USEPA Region 2 and the
RMW pertaining to evaluation of the potential human health risks associated with evaluation of the
suitability of dredged sediment for use as remediation material at the HARS.
In the sections below, each charge is provided as presented in the HARS framework, followed by the
consensus statement of the Panel and supporting information where necessary.
2.1 Response to Charge 1
Throughout the proposed process, there are various uncertainties introduced. Please identify the key
areas of uncertainty that need to be addressed. Are there additional data sources or parameters that
could be used to address these areas? What methods are available for describing and accounting for
these uncertainties in the calculation of HARS-Specific Values? Of the methods available, which would
you recommend for consideration and why? Please consider the implications of implementing these
methods in the regulatory framework. Please include an evaluation of probabilistic and deterministic
methods in your discussion.
Note: Responses to Charges 1, 16, 25, and 30 are addressed here.
Consensus Statement
The primary uncertainty identified by the Panel was associated with the relative lack of site-specific data
to support several key assumptions on which the HARS Framework was based. The Panel concurred that
the collection of additional data and the addition of full uncertainty and sensitivity analyses would
significantly improve the evaluation. The Panel agreed that probabilistic methods might be useful for
some portions of the evaluation, but acknowledged that sufficient data might not be available to use these
approaches for all of the models in the Framework. Among the most important uncertainties identified by
the Panel were: (a) characterization of contaminant levels in project sediment; (b) current environmental
conditions at the HARS (c) food web transfers of contaminants in sediment to prey and to upper trophic
level fish caught and consumed by anglers; and, (d) fishing and consumption habits among different
angler populations.
These four areas of uncertainty could be addressed by EPA either through field studies or modification of
one or more evaluation methods, including the use of probabilistic analysis. The Panel recognizes that
fully addressing data gaps associated with each of these areas of uncertainty will likely require, in some
cases, considerable effort over the next several years. Therefore, it was concluded that some components
of the proposed evaluation model might need to rely on a deterministic approach at least in the interim.
The Panel recommended that EPA consider the use of probabilistic methods to define specific parameters
(for which sufficient data are available) within the framework of the current deterministic model as an
interim approach until such time as new pertinent scientific information becomes available. However, the
Panel agreed that EPA needs to better justify the overall model structure for evaluating trophic transfer
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and selection of parameters, regardless of whether a deterministic or probabilistic approach to risk
assessment is used. A clear conceptual model should be constructed that critically examines the probable
food web structure occurring at the HARS. Once this is determined, the appropriate model type (i.e.,
deterministic or probabilistic) should be selected based on the availability of reliable data. The Panel
strongly recommended a sensitivity analysis and/or uncertainty analysis with whatever approach is
applied.
Justification
Uncertainty exists in all stages of the proposed process. Several resources are available describing the
numerous sources of uncertainty in human health risk assessment (e.g., USACE, 1999a; USEPA, 1989;
Paustenbach, 1995), sediment assessment (e.g., Ingersoll et al., 1997; Dickson et al., 1987), and sediment
bioaccumulation studies (e.g., Sharpe and Mackay, 2000).
To focus the evaluation of the various sources of uncertainty, the Panel identified three of the most
important aspects of the proposed evaluation where the process may be improved:
• Evaluation of background contamination;
• Trophic transfer; and,
• Human health exposure assessment.
Each of these topics is discussed below.
Evaluation of Background Contaminants
The Panel was not satisfied with the proposed approach for establishing reference bioaccumulation
potential for the region. The sandy reference site identified in the HARS Framework is inappropriate.
The Panel was unanimous that the proposed reference site possessed unique physical characteristics,
which will strongly affect bioavailability, and was wholly dissimilar from both the current conditions at
the HARS and from any proposed remediation material originating from NY/NJ Harbor. Specifically, the
proposed HARS Framework indicates that reference sediment is predominately (95%) sand, which
contrasts sharply with sediment conditions at the HARS and the physical characteristics of the majority of
project sediments in the NY/NJ Harbor region, which are predominately fine grained, organic rich silts
and clays. The toxicity and bioaccumulation potential of contaminants is generally considered to be
significantly lower in silty harbor and navigation channel sediments as compared to coarse-grained, sandy
organic-deficient materials such as encountered in the reference area. In addition the shortcomings of
using only one reference site, even one with similar physicochemical and biological characteristics poses
significant and unwarranted constraints on the proposed evaluation framework.
Consequently, it was the opinion of the Panel that more than one reference site encompassing the range of
background conditions in the NY Bight was needed. Alternatively, physical and chemical information
from several background reference areas could be compiled to develop a characterization of average
sediment conditions within the NY Bight. The Panel agreed that the EPA would greatly reduce
uncertainty in this phase by deriving a regional reference tissue level from empirical data. A similar
approach is applied by EPA Region 1, in which data from several potential reference sites were analyzed
to establish a range of chemical-specific tissue concentrations (Metcalf and Eddy, 1995). A percentile
ranking of background concentrations in the region was developed and threshold levels (e.g., 95%) were
established for comparison of bioaccumulation test data.
The Panel also addressed the possibility that dredged material might fail as suitable remediation material
based on consideration of the cumulative cancer and non-cancer risks to humans posed by exposure to
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contaminants in fish and shellfish. Under the current framework, dredged material could fail based on
evaluation of cumulative hazards and risks even if all chemical concentrations are below reference levels
in the event that background concentrations are elevated. The Panel encourages EPA to consider refining
the evaluation of cumulative risk to consider comparison to reference levels as well. Specifically, it was
suggested that the evaluation of cumulative risk should be conducted in comparison to cumulative risks
associated with regional, background concentrations, rather than an acceptable risk level (e.g., 10"4). One
reviewer cautioned, however, that this approach should be reconsidered if the cumulative risk associated
with reference was found to be significantly elevated over acceptable risk levels.
Recommendations:
• EPA Region 2 should develop regional reference tissue values based on a compilation of existing
data from monitoring programs and historical studies using the method employed by EPA Region
1. As an interim measure until regional values specific to the NY Bight can be developed, the
Panel recommended that the tissue values derived by EPA Region 1 (Metcalf and Eddy, 1995) be
applied.
• Dredged material should be determined suitable for use as remediation material if the cumulative
cancer and non-cancer risks to humans posed by exposure to contaminants in fish and shellfish
does not exceed the cumulative cancer and non-cancer health risks associated with the exposure
to contaminants in reference tissue levels, provided that reference concentrations themselves do
not pose a significantly elevated risk to human health.
Trophic Transfer
The Panel agreed that a better model for estimating trophic transfer should be established. They
recommended that the EPA document the trophic relationships (and associated uncertainties) linking
sediments to the selected fish (and shellfish) species. Ideally this would be done through stable isotope
analysis; within the various compartments of the food web, although in the absence of such information,
available data in the literature should be used to identify the appropriate links. It is possible that it will be
necessary to evaluate several distinct food webs to fully characterize exposure (e.g., consumption of
lobster, flounder, blue fish, etc.), however, the focus should be on those recreationally-important species
most likely to be highly exposed at the HARS (i.e., resident species) because they are the ones that will
pose the highest potential human health risks. Additional site-specific field data are needed to focus these
decisions and should be supplemented with literature-based approaches. Examples of potentially useful
investigations include evaluating tagging studies, modeling for each of the key species, evaluating
carrying capacity, or interpreting site productivity/life history data.
In this regard, the Panel defined resident species to mean organisms that 1) were likely to inhabit or spend
the majority of their life history at the HARS; and, 2) would be likely to be caught and consumed by
anglers fishing in the vicinity of the HARS. Invertebrates such as lobster and clams should be considered
in addition to fish. As defined, these resident species should represent the organisms that would
contribute the greatest risk in the human health evaluation.
In defining resident species, the Panel expressed concerns about using species that spend a large portion
of their time off-site and in areas that may be more heavily contaminated than the HARS. The Panel
suggested that if non-resident organisms (i.e., organisms that may visit the HARS but do not spend a
significant amount of time there) are also considered, only that portion of their exposure that is
specifically related to the HARS should be evaluated. The Panel suggested that the EPA should perform
a sensitivity analysis to determine if the risks associated with the consumption of these species are
significant relative to those species defined as resident prior to including them in the analysis. In general,
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the Panel felt that focusing on resident species would be more important for the trophic transfer
calculations.
Furthermore, the Panel observed that there is likely to be an inherent margin of safety associated with the
proposed evaluation process that has not been adequately described or recognized by EPA. Specifically,
if the HARS-specific Values are met for different contaminants, then it is likely that over time, surficial
sediment at the HARS will have average chemical concentrations that are significantly lower than the
regional threshold values. The Panel noted that EPA could confirm this and other environmental changes
at the HARS (as well as establish current conditions) through periodic monitoring over time.
The Panel agreed that EPA needs to better justify the overall model structure for evaluating trophic
transfer and selection of parameters, regardless of whether a deterministic or probabilistic approach to
risk assessment is used. A clear conceptual model should be constructed that critically examines the
probable food web structure occurring at the HARS. Once this is determined, the appropriate model type
(i.e., deterministic or probabilistic) should be selected based on the availability of reliable data. The Panel
strongly recommended a sensitivity analysis and/or uncertainty analysis with whatever approach is
applied.
Recommendations:
• Construct a clear conceptual model that defines the food web structure at the HARS site, focusing
on the most resident species. Verify food webs for critical species such as the lobster, flounder
and blue fish. This model should link sediments to potential human risks.
• Strengthen the evaluation with the collection of additional site-specific data and review of the
literature for species-specific information.
• Select the appropriate deterministic or probabilistic model (or elements of both) based on the
conceptual model and the data available.
• Clearly state the model assumptions and present a detailed uncertainty analysis.
Human Health Exposure Assessment
The Panel indicated that there was unacceptable uncertainty and ambiguity in many parts of the human
exposure assessment. The first issue that must be resolved is the development of a clear and precise
definition of the human receptor(s) to be evaluated and protected (e.g., the average or 95th percentile fish
consumer). Based on the information currently provided, it is not clear if the HARS-specific Values are
actually protective of a high-end fish and shellfish consumers within the exposed population. For
example, the data provided in NJMSC (1994) indicate that a recreational angler's average number of
meals per year could be obtained during only one fishing trip to the HARS. In addition, the dietary habits
of different ethnic groups differ significantly; it is possible that the consumption habits of some
recreational fisher's may include preparation of meals using the whole, gutted fish, and not just fillets as
assumed in the current approach. Additionally, shorter exposure durations are typically based on data
regarding residential mobility, however, when evaluating fishing behavior at an area such as the HARS,
residence in the region rather than in a single house or even town is the relevant consideration. As a
result, the proposed exposures may actually be underprotective for many anglers. EPA is encouraged to
address specifically whether the human health evaluation is intended to be protective of the average,
median, 90th percentile, or 95th percentile of the population of recreational fishers that may fish at the
HARS. EPA is also encouraged to specifically evaluate risks associated with short term exposures, thus
considering the results of multiple meals in a short time frame from one fishing trip to the HARS. This is
particularly an issue for sensitive life stages (e.g. fetal development).
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Once this is accomplished, site-specific data regarding fish consumption and the fraction of fish
consumption from the HARS can be developed. Variability and uncertainty in each of the parameters
need to be explicitly considered in the fish ingestion rate and other parameters. This could be done both
qualitatively within the context of a deterministic assessment or quantitatively within the context of a
probabilistic assessment.
Recommendations:
• Clearly define the receptor population to be protected;
• Focus attention on key species of fish taken recreationally from the HARS at the present time and
fully consider species that may be harvested in the future (e.g., lobster);
• Refine the consumption value based on site-specific information, the defined receptor population,
and the appropriate fish species;
• Revisit the assumption of consumption of only fillets rather than whole fish in the risk assessment
and consider a cooking loss factor (although it may appropriately be set to zero); and,
• Incorporate considerations of variability and uncertainty into the analysis.
2.2 Response to Charge 2
Is measurement of the 16 priority pollutant PAHs (i.e. parent PAHs) sufficient for characterizing the risks
associated with the total PAH bioaccumulated by organisms exposed to dredged material proposed for
placement at the HARS? Does measurement of the alkylated compounds significantly improve risk
assessment of PAHs?
NOTE: Charges 2,4, 5,13, 23,27, which address the potential toxicity of alkylated polycyclic aromatic
hydrocarbons (PAHs), were combined into a single consensus response to avoid overlap in answers to the
various questions. The issue of the approach for non-detects is addressed under the response to Charge
10.
Consensus Opinion
The Panel concurred that measurement of the 16 priority pollutant PAHs is not sufficient and that
alkylated PAHs should be included in the evaluation of human health risks. The primary uncertainty
introduced by inclusion of these chemicals was determined to be associated with the evaluation of their
toxicology, specifically the availability of human health benchmarks. However, given that many of the
alkylated PAHs are believed to be more toxic than the parent compound, it was concluded that the
resulting uncertainty was acceptable and that including these compounds would improve the analysis.
Justification
The Panel agreed that alkylated PAHs should be included in the assessment of PAH risk. While the Panel
recognized that alkylated PAHs are more associated with petrogenic sources of PAH than with pyrogenic
sources and, therefore, may be more prevalent at some locations than at others, it is anticipated that urban
influenced sites such as the NY/NJ Harbor would have sufficient petrogenic influence to warrant the
evaluation of the alkylated PAH for all sediments proposed as potential remediation material.
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With regard to the risk to humans posed by the alkylated PAHs, there is sufficient information in the
scientific literature to demonstrate that alkylated PAHs pose a risk, particularly for carcinogenicity. In
some cases, the health risks may be more serious than those associated with the parent compound. Thus,
omission of the alkylated PAHs in a risk assessment could potentially result in an underestimate of the
potential human health risks associated with dredged material.
The Panel recognizes that the toxicology, bioaccumulation and biotransformation of the alkylated PAH is
largely unknown. Clearly, the greatest source of uncertainty associated with inclusion of the alkylated
PAHs is the lack of human or animal toxicity data. To properly assess the toxicity of alkylated PAHs,
additional data are required. In the absence of specific data, the bioaccumulation potential should be
addressed on a log KoW basis with the selected log K<,w for each alkylated congener established either from
literature values or from a QSAR relationship. Quantitative Structure Activity Relationship (QSAR)
relationships for log K<,w are well recognized as providing accurate estimates of log KoW. It is
inappropriate to use the log Kow value of the parent PAH compound since the alkylated PAHs are more
hydrophobic. For biotransformation, the rate and extent of biotransformation should be related to PAHs
of similar log K«w unless specific information is available. For toxicology, where the data exist for
specific compounds, the specific human or animal data should be used for assessing risk. In most cases, it
will be necessary to assign potential toxicity based either on a QSAR relationship or, in the absence of
other information, using that of the parent compound. If the toxicity of the parent compound is used, it is
recognized that the potential risk will be underestimated for some mechanisms of action.
2.3 Response to Charge 3
Is the proposed adaptation of EPA Method 8270 (Appendix D) acceptable and appropriate for regulatory
decision-making? If not, what is an acceptable and appropriate method?
Consensus opinion
Based on the information presented, the Panel concluded that the analytical methodology known as EPA
8270, is acceptable and appropriate for determining concentrations of PAHs and related compounds and
for use in regulatory decision-making.
Justification
In general, the Panel felt that this methodology produced acceptable detection limits for PAHs and that
the quality assurance/quality control (QA/QC) procedures would ensure quality determinations. The
Panel noted that use of this analytical method would be advantageous because it is the same procedure
used in the National Oceanic and Atmospheric Administration's (NOAA) National Status and Trends
(NS&T) Program and other recent nation-wide surveys. However, there were a few concerns regarding
the protocol and associated procedures for controlling errors. For example, the data quality objective
(DQO) recovery targets should be 60 to 120 percent, not 40 to 120 percent as listed in Table 3. Also, the
analytical standardization of the method is based on the parent compounds (Table 2). It would seem more
appropriate to include some alkylated PAH compounds for surrogates, internal standards, and matrix
spikes to ensure that the method was performing well for the alkylated PAHs.
Even though the analytical methodology may be sound, there was also some concern that handling and
transport protocols be put in place to ensure that field samples are treated properly and that general
measures are taken to ensure that inadvertent or inappropriate contamination or cross contamination with
alkyl PAHs does not occur (e.g. diesel fumes from the boat). Field samples should also be stabilized
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immediately to avoid degradation and metabolism of the target analytes, especially the production of
alkyl PAHs from parent PAHs by microbial action.
An additional concern is that this method, as reported, does not include any N-heterocyclic PAHs on its
analyte list. Some of the N-heterocyclic compounds such as 9H-carbzole, 7H-dibenzocarbazole, and
acridine are known to be very mutagenic, and are usually not quantified in environmental samples
(Mastrangelo et al, 1996).
2.4 Response to Charge 4
Under what specific conditions would the testing for alkylated PAHs for a particular project be
appropriate and warranted?
Charges 2, 4, 5, 13, 23, 27 which address the potential toxicity of alkylated PAH were combined into a
single consensus response to avoid overlap in answers to the various questions. Please see the response to
Charge 2.
2.5 Response to Charge 5
What uncertainties would be introduced within the analysis of risk should alkylated PAHs be included?
What steps could be taken to account for these uncertainties in decision-making? Given the likelihood the
method for using non-detects (as described in EPA/CENAN, 1997) will result in an overestimate of risk,
what are the implications?
Charges 2,4, 5, 13, 23, 27 which address the potential toxicity of alkylated PAH were combined into a
single consensus response to avoid overlap in answers to the various questions. Please see the response to
Charge 2.
2.6 Response to Charge 6
It is recognized that additional methods have been used for the analysis of organotins (e.g., Krone et al,
1989). Will the proposed analytical method (Rice et al., 1987) provide adequate data of sufficient quality
to assess relevant risks from organotins? If not, please provide recommendations.
Consensus Opinion
The Panel concluded that the analytical method described by Krone et al. (1989) is the preferred approach
to measuring organotin compounds in dredged material.
Justification
The Krone et al. (1989) method was compared to the Rice et al. (1987) method and although they are
similar, some important differences were evident. First, Rice et al. (1987) was developed for tributyltin
rather than organotins and, unlike Krone et al. (1989), does not use tropolone, which is important for
extracting dibutyltin and monobutyltin. In addition, the Krone et al. (1989) method includes copper,
which is important in diminishing sulfur compounds that interfere with the flame photometric signal and
will considerably improve the detection limits.
Commercial laboratories that analyze for butyltins in environmental samples often use Krone et al.
(1989). Additionally, this method appears to address the important analytical considerations necessary to
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accurately assess the extent and impact of tributyltin (TBT)-related contamination in the marine
environment. These considerations include sample storage and preparation, accuracy and precision of
analytical methods, reporting convention, and quantitative method.
2.7 Response to Charge 7
What special QA/QC procedures should be implemented to ensure the quality and usability of the
organotin data?
Consensus Opinion
The Panel concluded that QA/QC procedures typically used for organotin analysis and described in
guidance from USEPA (2001) should be specified in the HARS Framework.
Justification
In general, improper sampling procedures, storage, manipulation, and analyses of environmental samples
can have significant impacts on chemical concentrations, speciation and bioavailability (USEPA 2001).
The Panel recommends that new guidance available from the USEPA (2001) should be closely followed
and documented to ensure proper QA/QC procedures are met. The potential for contamination of
reference samples, leading to the conclusion of insignificant metal enhancement in test samples is also
something that should be closely monitored. Although speciation changes or losses of organotins during
storage are not well established, a review of analytical procedures by Abalos et al. (1997) indicates that
freezing biological and sediment samples preserved the stability of organotins in the sample matrix for at
least 3 months.
For organotin analysis, appropriate recovery standards, gas chromatography (GC) internal standards and
certified reference material (CRM) are required. Additionally method blanks, spiked blanks, and
calibration standards also contribute to high quality analyses. For example, in the method of Krone et al.
(1989), tripentylmonobutyltin was employed as a GC internal standard. In this method, calibration was
done by the internal standard method using peak heights and a calibration curve with five concentrations.
For the recovery standard, tripentyltin chloride is commonly used. Additionally, TBT determinations
should be adjusted for the recovery of tripentyltin, which is run with each set. The limit of detection for
tissue and sediment should be close to 5 ng/g (dry weight).
For many years a tissue (sea bass) CRM has been available to gauge accuracy of tributyltin and
triphenyltin (Okamoto, 1991). The certified value is certified at 1.3 (0.1) /zg/g (mean and standard
deviation). Please see the individual response for Dr. James Meador (Appendix D) for information on
obtaining the seabass CRM. More recently, a sediment CRM (PACs II) has become available from the
National Research Council of Canada, which supplies a large number of CRMs for various chemicals and
matrices.
A related and important consideration is the reporting convention for butyltin concentrations. For
example, TBT is reported in different units, therefore, the numeric values calculated from these data are
often confusing to interpret. Specifically, TBT is reported as Sn, TBT, TBTC1, or TBTO (e.g., ng Sn/g or
ng TBT/g, or ng TBTO/g or molar). The most appropriate reporting unit is TBT ion, especially when
expressed in molar units.
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2.8 Response to Charge 8
Under what specific conditions would the testing for organotins for a particular project be appropriate
and warranted?
Consensus Opinion
The Panel recommends that every applicant should test for organotins in dredged material.
Justification
Most harbors and ports around the world are contaminated with organotins from shipping and dry-dock
activities. The Organotin Antifouling Paint Control Act of 1988 (OAPCA; US Congress 1988) banned
TBT paint on vessels less than 25 m, however, a large percentage of ships longer than 25 m are still using
TBT antifouling paint because it is so highly effective at controlling biofouling on ship hulls. Because
organotins are fairly persistent in sediment, elevated concentrations often occur in sediment in areas
where recreational craft predominate. Data summarized in Appendix E of the Peer Review Package
(EPA, 2001), although limited, indicate that average concentrations of (TBT) in polychaetes collected
from within the HARS are approximately twice that of polychaetes collected from the reference site.
These data suggest enrichment of TBT in sediments from the HARS. Moreover, data on concentrations
of TBT in surficial sediments taken from throughout the NY/NJ Harbor suggest that these compounds are
elevated throughout the area (Adams 1998).
Conceivably, at some sites, organotin concentration in sediment may be very low and thus, not pose a
concern for adverse effects. Therefore, it is suggested that applicants wishing to eliminate organotins
from their evaluations be allowed to submit additional information. Information to be considered could
include: 1) existing data on organotins in sediment to be dredged; 2) documentation of historic and
ongoing anthropogenic activities in the vicinity of the project; 3) sediment characteristics (e.g., organic
carbon content etc.); 4) arguments based on TBT partitioning in the environment (Meador, 2000); and 5)
the potential for biota to bioaccumulate organotins to levels that may be considered not harmful to
humans.
The Panel further recommends that EPA consider inclusion of phenyltins and immunotoxic organotins
that often occur in high concentrations in urban areas. Triphenyltin has been used in some antifouling
paints and as a fungicide in agricultural areas.
2.9 Response to Charge 9
If the approach for evaluating dioxin is modified, should it include the contribution of PCBs with dioxin-
like activity as proposed? If so, how?
Consensus Opinion
It was the consensus of the Panel that the co-planar polychlorinated biphenyls (PCBs) should be included
as part of the evaluation of dioxin-like activity for risk assessment. There is sufficient information
available to provide Toxicity Equivalent Factors (TEF) for the pertinent PCB congeners. Furthermore,
EPA's Draft Dioxin Reassessment (USEPA, 2000b) has concluded that including these compounds is
appropriate for risk assessment.
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Because of the testing scheme that evaluates PCBs and dioxin-like activity as separate pass/fail markers,
the co-planar PCB congeners should be included in assessments of both total PCBs and dioxin. However,
when the assessment proceeds to cumulative risk (e.g., total carcinogenicity), care should be exercised to
ensure that the health effects associated with exposure to co-planar PCB congeners are not double
counted.
2.10 Response to Charge 10
Please consider the policy for assigning values (at one half the detection limit) to tissue residues that are
reported as "
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2.11 Response to Charge 11
Is the use of functional groupings in statistical comparisons to reference appropriate and/or preferable to
statistical comparisons using individual contaminants for the purposes of risk analysis?
Consensus Opinion
The Panel recognizes that a long-standing issue in both human health and ecological risk assessments
concerns the appropriate methods for evaluating classes of chemicals that share similar physical and
chemical properties (e.g., dioxins, PCBs, PAHs, and DDT). The Panel agreed that it is useful to evaluate
functional groups of chemicals based on total molar concentration, but that often evaluation of individual
chemicals is more reliable due to a lack of scientific knowledge about functional grouping and a generally
accepted, broadly applicable classification scheme for classifying chemicals into appropriate groupings.
Therefore, the Panel concluded that the use of functional groupings for comparison of chemical
conditions in project and reference sediments was acceptable only where a clear rationale and justification
for the use of each functional grouping was provided. However, at least one peer reviewer strongly
recommended that the DDT compounds in particular not be included as a functional group, due to the
available information on the varying toxicity of the individual chemicals. In addition, at least one peer
reviewed noted that EPA mixture policy considers data on chemical mixtures preferential to combining
data on individual components (EPA, 1986).
Justification
In general, the Panel concluded that the use of functional groupings for comparison of chemical
conditions in project and reference sediments was acceptable where a clear rationale and justification for
the use of each functional grouping was provided. However, even currently employed functional
groupings are problematic. For example, with regard to the PCBs, available USEPA potency estimates are
predicated on total PCBs and selected Aroclor formulations and offer little insight into the behavior and
toxicity of the most important PCB congeners. Most scientists now agree that future PCB risk
assessments should evaluate non-ortho- and selected mono-ortho- PCBs, and use exposure models and
studies involving specific PCB congeners alone or in combination with the dioxins and furans (NRC,
2001; Eisler, 2000). For example, Coates and Elzermann (1986) demonstrated that equilibration times for
PCBs in sediments varied considerably from a few weeks for PCBs with low chlorine content to months
or years for PCBs with significantly higher chlorine content. Lick and Rapaka (1996) observed that slow
rates of adsorption / desorption may significantly modify the level of toxicity inferred from measurements
of the chemical concentration in the sediment, particularly if equilibrium partitioning is assumed but the
chemical in the pore water and colloidal phase is not in equilibrium with the chemical sorbed to the solid
phase. This will be a potentially important concern for aged PCBs in sediment, where degradation or a
change in the congener profile relative to the initial discharge may have changed substantially over time.
Similarly, PAHs as a general class of compounds, are not a functional grouping as they have subgroups
based on both chemical structure similarities and toxicological aspects; specifically, PAH subgroups
defined according to regulatory (priority) and chemical structure basis (alkyl PAHs) and according to
carcinogenic or noncarcinogenic effect endpoints. Similar to the PCBs, it may be more desirable in risk
assessment to evaluate individual PAH compounds rather than rely on the potency of one component
within a mixture or functional group. However, the state of the science may not be sufficient for
evaluating exposure to the individual analytes listed in Table 2 of the proposed HARS Framework.
Therefore, it may be necessary to group these compounds in order to evaluate exposures to the entire suite
of chemicals.
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The Panel also noted the HARS Framework does not clearly specify the appropriate procedure for
handling non-detect measurements in the course of calculating a functional group concentration. The
March 14,1997 EPA/CENAN document indicates that for total PCB and total DDT "conservative
estimates" of concentrations are to be used for constituents that are below detection limits; however, no
details on specific methods are presented. EPA is strongly encouraged to provide clarification.
Overall, the use of clearly defined functional groupings may be useful in risk assessment where there is
sufficient scientific/toxicological justification for the group and the use of such a grouping provides a
clear improvement in the assessment process by reducing uncertainty in the evaluation process.
Functional grouping for the HARS process may not necessarily be the same as schemes discussed in the
literature or employed elsewhere. It is recognized that for regulatory purposes, functional groupings
employed in the HARS evaluation process may not be completely scientifically sound but may be
required to facilitate decision making. Where this is done it should be explicitly noted that this is being
done to facilitate the decision-making process. Also, the possible uncertainties introduced by this process
and information that may help reduce these uncertainties should be noted.
2.12 Response to Charge 12
Is it appropriate to apply a multiplier based on log Kowfor these compounds (organics), or are there
other specific data that can be used to estimate steady state? If so, please identify.
Consensus Statement
The Panel concluded that, with the exception of organotins, the use of log K<,w as a multiplier for
estimating steady state for selected organic compounds (e.g., KoW < 6.0) is acceptable as an interim
approach. This approach will likely overestimate steady-state concentrations in benthos and is thus
appropriate given the regulatory context. However, the Panel recommended that regional-specific
multipliers based on synoptic measures of these compounds in sediments and benthos or chemical
specific multipliers based on laboratory evaluations be obtained to refine the more generic KoW based
multiplier. The Panel further recommended that the applicant be given the opportunity to provide
empirical data about steady state in place of the current multipliers until regional-specific multipliers are
available.
Justification
The Panel concurred that the use of K„w multipliers is a conservative approach because it does not take
into consideration metabolism of the compound, thus overestimating the final steady-state level. The
Panel concluded that, given the regulatory context, this approach is appropriate on an interim basis for
most chemicals with Kow values greater that 6.0.
The Panel cautioned, however, that this approach not be applied to organotins. Existing data suggest that
a steady state approach may not apply for organotins and that a KoW based multiplier would likely
underestimate organotins concentrations. EPA will need to develop an alternate method for organotins
based on the available data.
It was also noted that a number of technical issues add to the uncertainty of using K<,w multipliers. As an
example, this method does not account for biological factors that can affect bioaccumulation, time to
steady state, and steady state values, including temperature, body size, lipid content, gut surfactancy and
residence time, and rate of elimination (k2). Many of these factors are species-specific. Moreover,
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several Panel members questioned whether steady state conditions actually exist in the field given the
potential for significant temporal variations in exposure.
For these reasons the Panel recommended that the proposed approach be used only until it is possible to
develop either: 1) regionally specific multipliers based on synoptic measures of these compounds in
sediments and benthos; or 2) chemical specific multipliers based on laboratory evaluations of steady state
that include rates of elimination, since steady state depends in large part on elimination rate.
Finally, the Panel recommended that, while this policy is in place, the applicant be given the opportunity
to provide empirical data about steady state to reduce uncertainty and likely lower the estimated steady
state values.
2.13 Response to Charge 13
Given the increased hydrophobicity of alkylated PAHs, is the use of the correction factor associated with
the corresponding parent an appropriate approach for estimating steady state residues of alkylated
PAHs? If not, please elaborate.
Charges 2, 4, 5, 13, 23, 27 which address the potential toxicity of alkylated PAH were combined into a
single consensus response to avoid overlap in answers to the various questions. Please see the response to
Charge 2.
2.14 Response to Charge 14
For the DDT derivatives and dieldrin, please comment on the appropriateness of using M. nasuta data
rather than N. virens-specific data in the estimation of steady state multipliers.
Consensus Opinion
The Panel concluded that it is inappropriate to use M. nasuta data rather than N. virens data in the
estimation of steady state multipliers. The two organisms provide differing information about the
bioavailability of contaminants in the sediment and cannot be substituted one for the other. The Panel
encourages EPA to consider information for both species for the purposes of setting steady state
multipliers.
2.15 Response to Charge 15
Are the approaches taken to adjust organic contaminant bioaccumulation data to steady state adequate?
Do the proposed multipliers agree with previously published studies (i.e., do they appear reasonable)? If
not, please elaborate.
Consensus Opinion
The Panel concluded that, with the exception of organotins, the use of log Kow as a multiplier for
estimating steady state for selected organic compounds (e.g., K<,w < 6.0) is acceptable in the interim. See
the response to Charge 12 for more information.
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2.16 Response to Charge 16
What are the major sources of uncertainty associated with the approaches? What alternative approaches
would reduce the uncertainties? How could these uncertainties be described and accounted for in
decision-making ?
Charges 1, 16, 25, and 30 were combined. The major uncertainties associated with the human health
approach are summarized in response to Charge 1 and 18.
2.17 Response to Charge 17
In your opinion, is the methodology followed to derive the steady state multiplier for non-essential metals
(i.e., a factor of three) scientifically appropriate (Appendix X)? Please elaborate. Do you have any
recommendations of additional or alternate methodologies or information that can be used to either
supplement or replace the proposed method?
Consensus Opinion
EPA proposes to use a safety factor of three to adjust bioaccumulation data for metals from the 28-day
bioassay to a final 'steady state'. The Panel concluded that there was no scientific justification for the use
of this safety factor. The Panel recommended using available, synoptic field data from the New York
Bight to establish regional bioaccumulation factors and provide a baseline. In the absence of such
additional data, the Panel recommended that no safety factor be applied for any metal other than mercury.
The Panel recommended that a separate safety factor be developed for methylmercury, and that total
mercury be treated as methylmercury in the absence of data on mercury speciation.
Justification
EPA acknowledges that metals do not appear to reach a true steady state. Thus, they propose a safety
factor for non-essential metals based on the range of variability in concentrations of non-essential metal
(e.g., silver, cadmium, mercury and lead) in tissues of polychaetes from 14 stations in the New York
Bight. Because there is no more than a three-fold variation in concentration for any of the non-essential
metals, a safety factor of three was adopted for all of the metals. The approach is conservative in that it
assumes that the tissue data from the laboratory bioassays reflects the lowest concentrations likely to be
encountered in the field.
Safety factors are only proposed for non-essential metals "because there is less evidence for regulation of
non-essential metals by exposed marine organisms, and because of the higher level of human health and
ecological concern associated with non-essential (compared with essential) metals" (page 19 of the HARS
Framework). However, the very data relied upon for developing these safety factors suggests that benthic
organisms regulate accumulation of non-essential metals at least as effectively as they do essential metals.
Data in Appendix G of the HARS Framework show that polychaetes demonstrated less than three fold
variability in tissue concentrations of non-essential metals while synoptic sediment data varied by more
than 2 orders of magnitude, suggesting significant potential for regulating uptake. In addition, the range
of variability for non-essential metals (i.e., 2.3,2.8, 2.8 and 2.5 for silver, cadmium, mercury and lead
respectively) are consistent with those for the two essential metals evaluated in this study (i.e., 3.3 and 1.9
for copper and zinc, respectively). Therefore, the Panel concluded that these data do not support the
separate treatment of non-essential metals.
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Moreover, the Panel concluded that metals are generally not a bioaccumulation concern in aquatic food
webs, particularly in the marine environment, with the exception of methylmercury. Cadmium, lead and
nickel concentrations in tissue have been shown to decrease with increasing trophic levels and should thus
not be of concern for higher trophic level consumers such as humans. Based on these observations the
Panel concluded that, with the exception of mercury, the use of a safety factor for adjusting
concentrations of non-essential metals in tissues from bioassay results is not justified at this time for
evaluation of risk to humans.
In the case of mercury, the Panel was concerned that the proposed safety factor did not account for the
potential presence of methylmercury in sediments being considered as potential remediation material.
However, no specific recommendations are provided for addressing this issue.
2.18 Response to Charge 18
Please comment on each factor listed above (and in Table 5) as to its appropriateness for use in the
equations listed above. Would you recommend additional factors? Would you change or modify the
equations as written above? If so, how?
NOTE: Charge 18 and Charge 19 were addressed in a combined response.
Consensus Statement
The Panel agreed that the basic approach of using risk-specific doses to derive the HARS-Specific Values
was appropriate. The following modifications will result in redefinition of the parameter estimates for
fish intake, trophic transfer and site use factor. As a result, the current values specified in the proposed
HARS Framework for these factors will no longer be relevant:
• HARS-Specific Values should be set based on both acute and chronic risks. This will require two
separate analyses of exposure and risk. One analysis should be similar to the current calculation
and should focus on chronic effects that occur as the result of long-term exposures (i.e., greater
than one year) from fish consumption. The second analysis should focus on acute effects, in
particular reproductive and teratogenic effects. This assessment should use the same equations as
the first analysis but the fish ingestion rate should reflect the intake of fish that occurs over a short
period (i.e., 1-7 days). The predicted doses should be compared to relevant acute human health
benchmarks.
• The current approach of a site use factor is inadequate and should be replaced by a method that
links anglers' consumption of individual species and the spatial and temporal relationships
between the species and the contaminants in remediation material.
• The current equations for exposure should be modified to facilitate the use of probabilisitic
techniques to evaluate uncertainty and variability.
• The use of a fillet-to-whole fish factor may be inappropriate, and should be reconsidered on the
basis of site specific consumption information and consideration of future uses and angler
populations at the HARS. See response to Charge 20 for additional information.
• An additional factor for cooking loss should be included in the exposure assessment.
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In addition, the Panel expressed concerns with the scientific defensibility of several other exposure
factors, which will also be relevant to the revised equations. Concerns were raised regarding the
following factors:
• The basis for the fish consumption estimate is not appropriate and should be replaced with data
that are more relevant. The preferred source for such data would be a survey of peak and annual
average fish consumption rates by recreational marine anglers who fish the HARS or nearby
areas. Such a survey should collect data on the intake of specific species and the location where
the fish were harvested.
• The site use factor should be replaced with a factor that reflects the feeding range and prey of
each species harvested at the HARS.
• The duration of exposure should be based on the available demographic information of marine
anglers or from data collected in a new survey specifically focused on HARS anglers.
• Estimates of trophic transfer coefficients for specific contaminants are not adequately supported
by the available scientific evidence (see the response to Charge 1 for further discussion).
• The human health toxicity values used for certain contaminants may not reflect current
understandings.
Justification
The basic approach of establishing risk management goals (i.e., doses will not exceed the Reference Dose
or a dose associated with a 10"4 risk) is a reasonable approach for setting HARS-Specific Values.
However, the proposed system only evaluates chronic exposures and chronic health effects. Fish
consumption rates averaged over a year may be quite low (5-20 g/day); however, consumption of fish
actually occurs as a series of bolus doses (i.e., meals) that can result in daily intakes of 200-300 g/day. It
is also possible that a single fishing trip could result in multiple fish meals occurring over a short period
of time (one week). Many of the chemicals evaluated have acute reproductive and teratogenic effects that
may be caused by short-term peak doses. The system for setting HARS-Specific Values needs to assess
the potential for the contaminants to cause such acute effects from a single meal or a series of meals that
occur over a short period of time.
The proposed exposure equations use two factors, trophic transfer and site use factors, to characterize the
complex relationship between the proposed remediation material and exposed organisms. However, this
characterization fails to capture the spatial and temporal relationships between specific species and the
contaminants in the benthic organisms at the HARS. Each species will have a different relationship to the
residues in benthic organisms depending on the location where they are harvested and the season of the
year. These differences will be a function of the species' prey and feeding ranges. The location where a
species is caught will in turn determine the number and type of anglers who are exposed. The exposure
equations should be modified to capture these relationships.
As discussed elsewhere in these comments, EPA should give strong consideration to the use of
probabilistic models to explore the variability and uncertainty in the relationships between the HARS-
Specific Values and the doses received by recreational anglers. Using such models will require changing
the equations to define the intake, the individual's characteristics (body weight, cooking practices, and
duration of consumption). Ideally, probabilistic models would model the anglers' exposures on a fish by
fish basis, where the residue in each fish is modeled based on species-specific trophic transfer factors and
location where the fish was caught.
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The current basis for the fish consumption rate, the New Jersey Marine Sciences Consortium Study
(NJMSC, 1994), includes both marine and freshwater anglers. However, the productivity of freshwater
fisheries is generally less than marine fisheries, therefore, intakes of recreationally caught freshwater fish
will differ from marine fish. As a result, the current study is not appropriate for setting HARS-Specific
Values. In addition, it is not clear that the mean intake is the relevant measure of fish intake since this
would not protect high-end anglers. Anglers with fish consumption rates in the top 2-5% of the
distribution may have intakes that are 10 fold higher than the mean of the population. In addition, fish
consumption must be linked to the HARS site since the greatest risk will be associated with consumption
of fish that spend the most time at the HARS.
The preferred approach for improving the estimation of fish consumption is to obtain information from
anglers on the charter and privately owned fishing boats that actually fish at the HARS. These data can
be collected using intercept surveys of anglers at local marinas or by stationing boats at the HARS and
contacting the boats who fish the site. Alternatively telephone surveys of anglers in the counties nearest to
the HARS site could be performed.
The available demographic data (Marine Recreational Fishery Statistics Survey, 1991 and 1990; NJMSC
1994) indicate that most anglers in the region fish for less than 70 years. Anglers move into and out of
the State of New Jersey and many anglers take up fishing later in life. As a result, exposures for many
anglers may be less then a full lifetime. However, there is the possibility that lifelong anglers exist. For
this reason, a more careful evaluation of the available or site-specific demographic information should be
performed. In addition, the receptor of concern should be carefully identified in the conceptual site
model.
The trophic transfer factors proposed seem to be derived from a very limited number of laboratory
studies, and as some members of the Panel pointed out, seem lower than has been observed in actual
studies of trophic transfer in the environment. Rather than rely on a food web model where errors in
trophic transfer could dramatically over or under estimate risk, the panel suggests that empirical
measurements be made to establish food web position, and determine the transfer of model compound
classes in the HARS food chains. This could be accomplished through use of stable isotope analysis,
more complete evaluation of data available on trophic transfer of model contaminants in local species,
and limited laboratory experimentation to fill in data gaps.
Adopting a trophic transfer factor of 0.1 for PAHs seems very low. EPA does not reference the scientific
literature supporting their value of 0.1 for PAHs, except to say that it was obtained from "literature
values." Trophic transfer for benzo[a]pyrene and chlorinated compounds between algae and zebra
mussels has been reported as 0.53 and 0.95 (Bruner et al, 1999; Ma et al. 1999). Transfer between
sediment and benthic organisms has also been shown to exceed 0.1 by two to five-fold (Burner et al,
1999; Lydy and Landrum 1993). Even using 0.1 for the links to higher trophic levels where metabolism
of PAHs will be limiting may be an underestimate (McElroy and Sisson 1989; Varanasi 1989) in some
cases. Although significantly less bioavailable, the risks associated with trophic transfer of PAH
metabolites should not be ignored as they represent the predominant portion of the PAH-derived body
burden in benthos capable of metabolizing these compounds (McElroy and Sisson 1989), and can be
genotoxic themselves (McElroy et al, 1991). Furthermore, data on trophic transfer of alkylated PAHs is
very limited. The potential to underestimate trophic transfer using a factor as high as three exists for
chlorinated hydrocarbons where biomagnification in excess of a factor of three has been observed with
each trophic link (Rasmussen et al. 1990). At the very least, data such as these should be used to set the
limits on the range of predicted trophic transfer coefficients, and the uncertainty associated with this
variability explicitly included in the risk assessment. This represents one area where probabilistic
modeling approaches might be used to reduce uncertainty associated with a specific parameter.
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2.19 Response to Charge 19
Are the methods used to derive the human health exposure parameters and assigned values discussed in
Section E appropriate (please review the referenced appendices)? If not, please elaborate. How should
these factors be factored into the risk analyses and decision-making?
Consensus Statement
Please see Charge 18 for response.
2.20 Response to Charge 20
Is the approach taken to relate fish whole body and fillet concentrations scientifically appropriate? If
not, what method would you recommend?
Consensus Statement
The proposed approach specifies the use of generic, literature-derived ratios to predict fish fillet
concentrations from modeled whole-body concentrations of lipophilic compounds and selected metals.
These generic ratios are applied to all fish species. The Panel concluded that this approach did not
adequately account for the site-specific variability in whole body to fillet ratios that occur between
species and within species over time. The Panel recommended that site-specific whole body to fillet
ratios be developed based on data for those key species likely to be caught at the HARS and during the
seasons that these species are most likely to be caught. In the absence of this site-specific data, the whole
body/fillet ratio should be dropped from the exposure calculation (or a whole body/fillet ratio of 1
substituted for the existing ratios).
Justification
Appendix K of the HARS Framework proposes that a single whole-body to fillet ratio of 1.35 be used for
all lipophilic compounds and all fish species. At steady state, lipophilic compounds will partition among
tissues of the fish based on the lipid content of those tissues. The proposed approach assumes that the
lipid content of the fillet and whole body are consistent among fish species. However, the relative lipid
content of tissues can vary substantially among fish species and over time within species. Appendix K of
the HARS Framework indicates that proposed ratio was derived from fish from New York State and the
Great Lakes. It is unclear which species were used in deriving the proposed ratio or how lipid content of
the tissues of these species might compare with those that forage in the area of the HARS.
For inorganic compounds, whole-body to fillet ratios for arsenic, chromium and mercury are derived from
regression analysis of data from black bass (Bevelhimer et al., 1997). The underlying assumption in
using these ratios is that the species and exposure range from this study are representative of the species
of concern and exposure ranges found at the HARS. However, in the absence of site-specific data it is not
possible to evaluate this assumption. This was of particular concern for mercury, as the whole-body to
fillet ratio may vary substantially depending on the proportions of methylmercury and inorganic mercury
in fish tissues. A whole-body to fillet ratio of one is used for the remaining metals (Ag, Cd, Cu, Ni, Pb
and Zn); however, no justification is presented. Although a ratio of one is likely to considerably
overestimate concentrations of metals in fillet, it does not accurately reflect the relationship between the
concentrations in the whole-body and fillet as fish are able to regulate tissue-level compartmentalization
of metals such as Cd, Cu, Pb and Zn over a wide range of exposures.
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Given this variability, the Panel recommended that site-specific whole body to fillet ratios be developed
for those key species likely to be caught at the HARS. This evaluation should be conducted during the
seasons that these species are most likely to be caught. Site-specific data on the distributions of whole-
body to fillet ratios for each class of chemicals would provide insight as to the variability and uncertainty
associated with this parameter and thus a more scientifically defensible basis for estimating exposure via
this pathway.
In the absence of site-specific data, the whole body/fillet ratio should be dropped from the exposure
calculation, or a whole body/fillet ratio of one substituted for the existing ratios. A whole-body to fillet
ratio of one likely overestimates concentrations of both lipophilic and inorganic chemicals in fillets and
thus provides a protective alternative for evaluating risk in the absence of site-specific data.
2.21 Response to Charge 21
Could the analysis be improved by focusing on key fish (seafood) species at the HARS? What
characteristics should be used to select these key species?
Consensus Opinion
The Panel felt strongly that the EPA should focus on key seafood (i.e., fish and shellfish) species actually
foraging and caught at the HARS. In addition, they should use bioaccumulation and food chain transfer
data determined for these species in the determination of health risks to anglers.
Justification
The Panel recommended that species included in the model for human health risk assessment should be
those caught as part of the recreational fishery both at the present time and that may be included in the
future. In addition, the risk assessment should focus on those species resident in the HARS for a
significant portion of their life cycle. The best species to target would be those that live in close
association with the sediment, have high lipid contents, and limited ability to metabolize contaminants.
Species that forage for a significant portion of their lives in other areas should not be included in the
analysis unless they represent a significant portion of the catch from the HARS. If included, their
contribution to the HARS-specific risk should be calculated based on the percentage of their contaminant
body burden that can be attributed to exposures from the HARS. Initial efforts should be focused on
species caught by recreational anglers, but species captured as part of the commercial fisheries should not
be ignored.
Using these criteria, the approach should focus on summer and winter flounder, tautog, and black sea bass
as predominately resident species, and bluefish as migratory species spending a significant portion of
their life history in the HARS. In addition, lobsters and a molluscan species, either hard clams or
scallops, should also be included as representative of the HARS commercial fishery. There is a wealth of
information available on the life histories, migration patterns, trophic relationships, and in some cases,
contaminant body burdens of these species. EPA would minimize uncertainty in assessing risk if it
focused on the species resident at the HARS that are most frequently captured by local recreational and
commercial anglers. By estimating risk associated with consumption of these species, the range of
potential dietary exposure to local anglers and consumers of locally caught seafood would be captured.
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Recommendations:
• Risk assessment should focus on key seafood species resident at the HARS that spend a
significant portion of their life cycle at the HARS
• Species selected should be important to the fisheries, reside primarily at the HARS, live in close
association with the sediment, have high lipid content, and limited ability to metabolize
contaminants.
• Species that spend a significant portion of their time foraging in other locations should only be
included, if they represent a significant portion of the HARS fishery, and then only if the non-
HARS contaminant contribution can be estimated.
• Both fish and shellfish (crustacean and molluscan) species should be included in the assessment.
2.22 Response to Charge 22
In your opinion, is the approach for assuming total metal to be in the most toxic form appropriate and
reasonable? Should metal speciation/complexation be considered in the assessment of metals
bioaccumulation, trophic transfer, and human health risks? Is the proposed approach for evaluating
methyl mercury appropriate? Are there alternative analytical or risk assessment techniques available that
would improve the risk assessment of metals? Is the multiplier proposed for adjusting measured
concentrations of arsenic appropriate and reasonable?
Consensus Statement
The Panel agreed that, in the absence of data on metal speciation, it is appropriate to assume that the total
metal concentration is equivalent to the most toxic form or species of each metal. This approach is
acceptable to the Panel as long as the EPA utilizes appropriate project-specific data on metal speciation
when made available by the applicant.
Justification
For chromium and mercury, the use of potency and exposure factors for the most toxic valence species or
form of the metal will most likely overestimate risk in a marine environment such as the HARS. This
conservative approach is justified given the uncertainty resulting in the absence of data on speciation of
chromium and mercury. Moreover, the policy allows proponents of dredging projects to provide data on
metal speciation to reduce this uncertainty.
For arsenic, the use of a factor of 0.1 will likely overstate the actual concentrations of inorganic arsenic in
bioassay species. Again, this conservatism is appropriate in the absence of data on arsenic speciation.
However, as with chromium and mercury, proponents of dredging should be given the opportunity of
providing data on arsenic speciation rather than relying on a generic application factor.
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2.23 Response to Charge 23
Is the assumption that the potency of alkylated PAHs can be estimated by the potency of the parent PAH
appropriate? Is this assumption likely to result in an under- or overestimate of the risk associated with
the alkylated PAHs?
Charges 2,4, 5, 13, 23, 27 which address the potential toxicity of alkylated PAH were combined into a
single consensus response to avoid overlap in answers to the various questions. Please see the response to
Charge 2.
2.24 Response to Charge 24
Please comment on the potential for human exposure to PAHs through consumption offinfish and other
seafood.
Consensus Opinion
It was the consensus of the Panel that human exposure to parent PAHs through consumption of finfish
would likely be minimal due to the biotransformation capability of fish. However, this is not the case for
consumption of organisms such as clams and even lobster that have lower capabilities for
biotransformation. The biotransformation potential of each of the food webs under consideration for
human consumption should be evaluated for biotransformation potential to assess the extent of exposure
to parent PAH. It was also noted that there would likely be exposure to PAH metabolites that may be
toxicologically active. However, the toxicity potential of metabolites is unknown.
2.25 Response to Charge 25
What are the major sources of uncertainty associated with the approaches described in Section E? What
alternative approaches would reduce the uncertainties? How could these uncertainties be described and
accounted for in decision-making?
Charges 1,16,25, and 30 were combined. The major uncertainties associated with the human health
approach are summarized in response to Charge 1 and 18.
2.26 Response to Charge 26
What is your recommendation for evaluating the potential toxicity of organotins? Should they be
evaluated as individual compounds? Summed as total? Should there be some consideration of relative
toxicity?
Consensus Opinion
The Panel concluded that organotins should be evaluated collectively, with a focus on TBT.
Justification
Human health risks from butyltins is mostly associated with tributyltin, which has been shown to be toxic
to the thymus-dependent immune system. The EPA has set a reference dose for TBT at 0.3 ug/kg/day;
however, there are other published values ranging from 0.25 to 6.7 ug/kg/day (see Cardwell et al. 1999
for a review). Dibutyltin has also been shown to be thymotoxic in laboratory studies (Seinen 1980), is
known to inhibit oxidative phosphorylation and is generally cytotoxic (see review by Snoeij et al. 1987).
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No toxicity data were found for monobutyltin, so it is not known if this compound will produce similar
toxicity in humans. Tetrabutyltin, if it occurs, should also be included because it will be metabolized to
tri- and dibutyltin.
The Panel recommended treating tetrabutyltin, tributyltin, and dibutyltin together in a human health risk
assessment. One approach is to consider these compounds equally toxic, which would allow use of the
RfD set for TBT. Additionally, it may be possible to derive the RfD for dibutyltin from the literature and
apply it to the observed data.
2.27 Response to Charge 27
Comment on the appropriateness of the proposed approach for converting and using the analytical data
for alkylated and parent PAHs to estimate risk from all PAHs.
Charges 2, 4, 5, 13, 23, 27 which address the potential toxicity of alkylated PAH were combined into a
single consensus response to avoid overlap in answers to the various questions. Please see the response to
Charge 2.
2.28 Response to Charge 28
Do you believe that the "disaggregate" modeling discussed above (and shown in Figure 3) for estimating
human health HARS-Specific Values for lead is appropriate? Would you recommend an alternative risk
assessment method be used given the information and data available? Do you believe the method
described has appropriately taken uncertainty into account? Please elaborate.
Consensus Opinion
The Panel concurred that the application of a disaggregate modeling approach for evaluating exposures to
lead is appropriate. However, it was noted that approach typically used in human health risk assessment
to evaluate lead exposure in adults and children is the integrated exposure uptake/biokinetic (IEUBK)
model (USEPA, 1999b, 1994a and b). The approach used in the proposed HARS Framework follows a
slightly different approach based on an older framework (EPA, 1986a and b) and should be evaluated
closely with regard to the biokinetic approach. In addition, edits to specific parameters within the model
should be included as discussed in response to other charges (e.g., fish ingestion rate, whole body to fillet
ratios, etc.). The Panel also noted that the disaggregate modeling approach for lead typically focuses on
exposures to children, however, the ingestion rate proposed was developed for adults. If EPA proceeds
with the approach as currently described, the Panel suggests that information on background exposure
levels likely encountered in the NY/NJ metropolitan area be incorporated as well as consideration of these
other parameters.
2.29 Response to Charge 29
In your opinion, are the methodologies and equations described above appropriate for estimating total
carcinogenicity and combined non-cancer impacts of contaminant mixtures accumulated from dredged
materials proposed for use as Remediation Material at the HARS?
Consensus Opinion
The equations and methodologies for characterizing the total carcinogenicity of contaminant mixtures in
proposed remediation material are generally consistent with widely applied and accepted risk assessment
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methodology and are, therefore, suitable. However, the cumulative non-cancer impacts should be
evaluated for all contaminants, not just those listed in Table 6 of the HARS Framework, to provide a
comprehensive assessment of the non-cancer effects. Carcinogenic compounds also have non-
carcinogenic effects and should be included in the consideration of cumulative impacts. For example, the
absence of PCBs, mercury, and many PAHs from the summary of contaminants/effects is particularly
notable. The Panel also concluded that the list of non-cancer effects of concern should be expanded to
consider, where possible, reproductive, developmental, endocrine, and immunological effects. As basic
toxicological information and detailed regulatory analysis (e.g., non-cancer total daily intake (TDI) or
reference dose (RfD) for non-cancer endpoints) are not available for some chemicals, there will be gaps
and limitations in assessment of additional non-cancer effects until appropriate information becomes
available.
Justification
The overall approach for evaluating the combined effects of multiple contaminants, i.e., evaluating the
individual contaminants at a specified risk level, and then evaluating the combination for the same risk
level, is acceptable, although it differs from that used in some other EPA programs. For example, for the
preliminary remediation goals used in Superfund, a lower risk level (frequently 10-fold) is used for the
individual compounds than what is intended for the total site risk. However, the approach proposed for
combining effects should be fully protective if the summed effect risk value is not allowed to exceed the
benchmark risk value. The proposed approach has the advantage of allowing for a large amount of
variability among different dredged materials. It allows for materials with relatively high concentrations
of a single contaminant (and low concentrations of other contaminants) and materials with moderate
concentrations of multiple contaminants to be evaluated against the same, final risk standard.
The details of the proposed approach for cancer risk are appropriate as provided in the document. The
assumption that the cancer risks from the individual compounds are additive is commonly used as the
default regulatory assumption, although it is recognized as a simplification based on limited
understanding complex mixtures. What is likely to be the most difficult part of the cancer risk summation
approach is not addressed in the document, namely a method to address the risk of PCBs and the dioxin
toxic equivalence (TEQ) associated with individual PCB congeners such as 126 and 169. This
onundrum will arise if (or when) the approach includes tetrachlordibenzo-p-dioxin (TCDD) and dioxin-
like compounds in the risk based approach instead of the pass/fail concentration criteria that now exist.
The approach for evaluation of non-cancer effects should be more comprehensive. It should include the
non-cancer effects of all contaminants, not just those not considered carcinogenic. This is not "double-
/ counting" the effects of the contaminants, as many of the carcinogens are toxic in multiple modes and
Jhus contribute to the total toxic risk in multiple organ systems.
The non-cancer effects that are included in the evaluation should be more comprehensive. Many of the
contaminants with high bioaccumulation potential, such as PCBs, DDT, and mercury, are associated with
/l/Endocrine, reproductive, developmental and/or immunological effects. If TCDD and dioxin-like
. f rAcompounds are evaluated with respect to risk, these will also contribute substantially to these non-cancer
vC ^effects. The inclusion of these non-cancer effects in the evaluation will not be a trivial exercise, however,
V\l as the RfDs may be based on other effects (e.g., the RfD for DDT is based on liver toxicity), and some
r consideration of the relative doses at which the different effects are observed may be necessary.
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2.30 Response to Charge 30
Is the conceptual model for evaluating fish exposure to dredged material at the HARS and human
exposure through ingestion of seafood appropriate and reasonable? How can the uncertainties
associated with the assumptions in this conceptual model be reduced? Please consider the spatial and
temporal elements of exposure in your discussion.
Charges 1, 16, 25, and 30 were combined. The major uncertainties associated with the human health
approach are summarized in response to Charge 1 and 18.
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3.0 REFERENCES
Abalos, M., Bayona, J., Compano, R., Granados, M., Leal, C., and Prat, M. 1997. Analytical procedures
for the determination oforganotin compounds in sediment and biota: critical review. Journal of
Chromatography A, 788 (1-49)
Adams, D.S., J.S. O'Connor, and S. B. Weisberg. 1998. Sediment Quality of the NY/NJ Harbor System.
An Investigation under the Regional Environmental Monitoring and Assessment Program (R-EMAP).
Final Report. EPA/902-R-98-001.
Bevelhimer, M.S., JJ. Beauchamp, B.E. Sample, and G.R. Southworth. 1997. Estimation of whole-fish
contaminant concentrations from fish fillet data. U.S. Department of Energy, ES/ER/TM-202.
Bruner, K.A., Fisher, S.W. and Landrum, P.F. 1999. The role of the zebra mussel, Dreissena polymorpha,
in contaminant cycling: II. Zebra mussel contaminant accumulation from algae and suspended particles
and transfer to the benthic invertebrate, Gammarus fasciatus. J. Great Lakes Res. 20:735-750.
Cardwell, R,D, Brancato, MS., Toll, J., DeForest, D., Tear, L. 1999. Aquatic Ecological risks posed by
tributyltin in University States surface water: pre-1989 to 1996 data. Env. Toxicol, and Chem. 14:567-
577.
Coates, J.T., Elzermann, A.W. 1986. Desorption kinetics for selected PCB congeners from river
sediments. J. Contam. Hydrol. 1:191-210.
Dickson, K.L., Maki, A.W., Brungs, W.A. 1987. Fate and Effects of Sediment-Bound Chemicals in
Aquatic Environments. Pergamon Press, New York, NY. pp. 449.
Eisler, R. 2000. Handbook of Chemical Risk Assessment. Lewis Publishers, Boca Raton, FL. Chapter 25.
EPA (see also U.S. Environmental Protection Agency).
EPA/CENAN (U.S. Environmental Protection Agency, Region 2/U.S. Army Corps of Engineers, New
York District) 1997. (Joint Memorandum) Ocean Disposal of Dredged Material Clarification of Two
Procedural Elements of Interagency Coordination Between USEPA Region 2 and the New York District,
USACE-Treatment of Non-Detects, Chemical Data, and Rules and Responsibilities in Preparation of
Ocean Disposal Regulatory Compliance Memorandum.
Ingersoll, C.G., Dillon, T., Biddinger, R.G. 1997. Methodological uncertainty in sediment ecological risk
assessment. In: Ingersoll, C.G., Dillon, T., Biddinger, G.R. (Eds.). Ecological Risk Assessments of
Contaminated Sediment. SETAC Press, Pensacola, Florida, 389 pp.
Krone, C., Brown, A.D.W., Burrows, D.G., Bogar, R.G., Chan, S-L., Varanasi, U. 1989. A method for
analysis and measurement ofbutyltins in sediment and English sole livers from Puget Sound. Mar.
Environ. Res. 27:1-18.
Lick, W, Rapaka, V. 1996. A quantitative analysis of the dynamics of the sorption of hydrophobic organic
chemicals to suspended sediments. Environ. Toxicol. Chem. 15(7): 1038-1048.
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Lydy, M.J. and Landrum, P.F. 1993. Assimilation efficiency for sediment-sorbed benzo[a]pyrene by
Diporeia spp. Aquat. Toxicol. 26:209-224.
Ma, X., Bruner, K.A., Fisher, S.W., and Landrum, P.F. 1999. Absorption of hydrophobic contaminants
from ingested chlamydomonas rheinhardtii and chlorella vulgaris by zebra mussels, Dreissena
polymorpha. J. Great Lakes Res. 25:305-317.
Marine Recreational Fishery Statistics Survey, 1991 and 1990.
(http://www.psmfc.org/recfin/mrfssov.htmI
Mastrangelo, G., Fadda, E., and Marzia, V. 1996. Polycyclic aromatic hydrocarbons and cancer in man.
Environ. Health Perspect. 104:1166-1170,
McElroy, A.E. and Sisson, J.D. 1989. Trophic transfer of benzo(a)pyrene metabolites between benthic
organisms. Mar. Environ. Res. 28:265-269.
McElroy, A.E., Cahill, J.M., Sisson, J.D., and K.M. Kleinow. 1991. Relative bioavailability
and DNA adduct formation of benzolajpyrene and metabolites in the diet of the winter flounder. Comp.
Biochem. Physiol. 100C:29-32.
Meador, J.P. 2000. Predicting the fate and effects oftributyl tin in marine systems. Rev. Environ.
Contam.Tocicol. 166: 1-48.
Metcalf and Eddy. 1995. Background concentrations of contaminants in benthic invertebrate tissue, US
Environmental Protection Agency, Region 1.
National Research Council (NRC). 2001. A Risk Management Strategy for PCB-Contaminated Sediments,
National Academy Press, Washington, D.C. January.
New Jersey Marine Sciences Consortium (NJMSC). 1994. Fish Consumption Patterns by New Jersey
Consumers and Anglers. Prepared for the New Jersey Department of Environmental Protection and
Energy, Division of Science and Research.
Newman, M.C., P.M. Dixon, B.B. Looney, and J.E. Pinder III, 1989. Estimating mean and variance for
environmental samples with below detection limit observations. Water Res. Bull. 25:905-916.
NY/NJ HEP. 1996. Final Comprehensive Conservation and Management Plan. New York-New Jersey
Harbor Estuary Program, Including the Bight Restoration Plan. March 1996
Okamato, K. (1991) Biological reference materials for meted speciation: NIES fish tissue reference
material for organotin compounds. In K.S. Subramanian, G.V. Iyengar and K. Okamoto (eds.)
Biological Trace Element Research: Multidisciplinary Perspectives. American Chemical Society
Symposium Series no. 445. Washington, D.C. pp. 257-264.
Paustenbach, D.J. 1995. Retrospective on U.S. health risk assessment: How others can benefit. Risk
Health Saf. Environ. (6):283-332.
Rasmussen, J.B., Rowan, D.J., Lean, D.R.S., and Carey, J.H. 1990. Food chain structure in Ontario lakes
determines PCB levels in lake trout (Salvelinus namaycush) and other pelagic fish. Can. J. Gish. Aquat.
Sci. 47:2030-2038.
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Rice, C., F. Espourteille and R. Huggett. 1987. A method for analysis of tributyltin in estuarial sediments
and oyster tissue, Crassostrea virginica. Appl. Organomet. Chem. 1:541-544.
Seinen, W. (1980). Immunosuppression induced by certain organotin compounds. Vet. Sci. Commun.
3:279-287.
Shaipe, S., Mackay, D. 2000. A framework for evaluating bioaccumulation in food webs. Environ. Sci.
Technol. 34(12):2373-2379.
Snoeij, N.J., A.H. Penninks, and W. Seinen. (1987). Biological activity of organotin compounds-an
overview. Environmental Research 44:335-353.
US Congress (1988). Organotin antifouling paint control act of 1988. HR 2210-3. One Hundredth
Congress of the United States of America, Second Session. Washington, DC.
U.S. Environmental Protection Agency (USEPA). 1986a. Air Quality Criteria for Lead. Research
Triangle Park, NC: Office of Research and Development, EPA 600/8-83-028F.
U.S. Environmental Protection Agency (USEPA). 1986b. Guidelines for the Health Risk Assessment of
Chemical Mixtures. Federal Register 51FR34014-34025.
U.S. Environmental Protection Agency (USEPA). 1989. Risk Assessment Guidance for Superfund. Vol.
1, Human Health Evaluation Manual (Part A). EPA/540/1-89/002. Office of Emergency and Remedial
Response. Washington, D.C. December.
U.S. Environmental Protection Agency (USEPA). 1994a. Guidance Manual for the Integrated Exposure
Uptake Biokinetic Model for Lead in Children. Washington, D.C. Office of Solid Waste and Emergency
Response. EPA/540/R-94/039.
U.S. Environmental Protection Agency (USEPA). 1994b. Technical Support Document: Parameters and
Equations used in Integrated Exposure Uptake Biokinetic Model for Lead in Children. Office of Solid
Waste and Emergency Response. Washington, D.C. EPA/540/R-94/040.
U.S. Environmental Protection Agency (USEPA). 1999a. Risk Assessment Guidance for Superfund:
Volume 3- (Part A, Process for Conducting Probabilistic Risk Assessment). Draft. EPA/000/0-99/000.
Office of Solid Waste and Emergency Response, Washington, D.C. December.
U.S. Environmental Protection Agency (USEPA). 1999b
www.epa.gov/superfund/programs/lead/products/rmieubk32.pdf
U.S. Environmental Protection Agency (USEPA). 2000a. Science Policy Council Handbook - Peer
Review 2nd edition. EPA 100-B-00-001. Office of Science Policy. Washington, D.C. December 2000.
U.S. Environmental Protection Agency (USEPA). 2000b. Methods for Measuring the Toxicity and
Bioaccumulation of Sediment-associated Contaminants with Freshwater Invertebrates - Second Edition.
EPA-823-F-00-002, U.S. Environmental Protection Agency, Washington, DC.
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U.S. Environmental Protection Agency (USEPA). 2001. Scientific Peer Review Package and Charge:
Proposed Bioaccumulation Testing Evaluation Framework for Assessing the Suitability of Dredged
Material to be Placed at the Historic Area Remediation Site (HARS). Draft report prepared by the United
States Environmental Protection Agency. December 21, 2001.
Varanasi, U., J.E. Stein, and M. Nishimoto. 1989. Biotransformation and disposition of polycyclic
aromatic hydrocarbons (PAH) in fish. In: Varanasi U. (ed) Metabolism of Polycyclic Aromatic
Hydrocarbons in the Aquatic Environment. CRC Press, Boca Raton, FL, pp 94-149.
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Figure 2-1. Proposed Peer Review Process
PHASE I
HUMAN HEALTH
PHASE II
ECOLOGICAL EFFECTS
RMW Meeting: Human Health &
Shared Issues Charge Input
(2 Days)
RMW Meetings:
Ecological TEF Development
60 Days
f 30 Days
Peer Reviewer Meeting: Overall Briefing I
and Human Health & Shared Issues I
Charge Delivery (2 Days) I
T ^
Prepare Ecological TEF
for Example Compounds
Independent Scientific Review of
Health & Shared Issues
Human I
120 Days
RMW Meeting: Ecological Issues
Charge Input
60 Davs
Peer Review Meeting: Issue Resolution
for Human Health & Shared Issues
(1 iw^
30 Davs
Peer Reviewer Meeting: Ecological
Issues Charge Delivery
Human Health & Shared Issues
Interim Report
Independent Scientific Review of
Ecological Issues
60 Davs
Peer Reviewer Meeting: Issue
Resolution for Ecological Issues
(2 Days)
Interim Report: Ecological Issues
(30 Days)
Human and Ecological
Consensus Report
I
RMW
Briefing
—r
Prepare Draft TEF
(3 Months)
V
Review/Finalize Responsiveness
Report
(2 Months)
I
RMW Review and Discussions
Revise Draft TEF
(21 days)
Q
RMW Only
Q
Peer Reviewers Only
Q
RMW & Peer
Reviewers
Public Comment
(60 Days)
Finalize Technical Resolution and
TEF Amendments
Finalize TEF
-------
Appendix A
Final Panel for HARS Peer Review
-------
Final Panel for HARS Peer Review
Dr. Allen Burton
Wright State University
Institute for Environmental Quality
064 Brehm Laboratory
3640 Colonel Glenn Hwy
Dayton, Ohio 45435-0001
937-775-2201
allen.burton@wright.edu
Dr. Ken Jenkins
Arcadis JSA
301 East Ocean Blvde, Suite 1530
Long Beach, CA 90802
562-628-1176
kj enkins@arcadis-us. com
Dr. Peter Landrum
US Department of Commerce, NOAA
Great Lakes Environmental Research
Laboratory
2205 Commonwealth Blvd
Ann Arbor, MI 48105
313-741-2276
landrum@glerl.noaa.gov
Dr. Lynn McCarty
94 Oakhaven Drive
Markham, Ontario L6C 1X8
CANADA
905-887-0772
lsmccarty@rogers.com
Dr. Anne McElroy
Marine Sciences Research Center
State University of New York
Stony Brook, NY 11794-5000
631-632-8488
amcelroy@notes.cc.sunysb.edu
Dr. Jim Meador
Environmental Conservation Division
Northwest Fisheries Science Center
2725 Montlake Blvd
Seattle, WA 98112
206-860-3321
j ames.meador@noaa.gov
Mr. Paul Price
129 Oakhurst Road
Cape Elizabeth, ME, 04107
207-799-3406
psprice@pipeline.com
Dr. Harlee Strauss
21 Bay State Road
Natick, MA 01760
508-651-8784
h.strauss@rcn.com
Mr. Rick Wenning
Environ Corp
Marketplace Tower
6001 Shellmound Street, Suite 700
Emeryville, CA 94608
510-655-7400
rjwenning@environcorp.com
-------
G. Allen Burton, Jr., Ph.D.
Dr. G. Allen Burton, Jr. is the Brage Golding Distinguished Professor of Research and Director
of the Institute for Environmental Quality and Ph.D. Program at Wright State University. He
obtained a Ph.D. degree in Environmental Science from the University of Texas at Dallas in
1984. From 1980 until 1985 he was a Life Scientist with the U.S. Environmental Protection
Agency. He was a Postdoctoral Fellow at the National Oceanic and Atmospheric
Administration's Cooperative Institute for Research in Environmental Sciences at the University
of Colorado. Since then he has had positions as a NATO Senior Research Fellow in Portugal and
Visiting Senior Scientist in Italy and New Zealand. Dr. Burton's research during the past 20
years has focused on developing effective methods for identifying significant effects and
stressors in aquatic systems where sediment and stormwater contamination is a concern. His
ecosystem risk assessments have evaluated multiple levels of biological organization, ranging
from microbial to amphibian effects. He has been active in the development and standardization
of toxicity methods for the U.S. EPA, American Society for Testing and Materials (ASTM),
Environment Canada, and the Organization of Economic Cooperation and Development (OECD).
Dr. Burton has served on numerous national and international scientific committees and review
panels, has had over $4 million in grants and contracts, and over 100 publications dealing with
aquatic systems.
-------
Kenneth D. Jenkins, Ph.D.
Dr. Jenkins has over 30 years of experience in the field of environmental toxicology. He is a
Senior Vice President of Blasland, Bouck & Lee (BB&L). He served as a Professor of Biology at
California State University at Long Beach (CSULB) from 1970 to 1997 and is now a Professor
Emeritus. While at CSULB he founded and directed the Molecular Ecology Institute and taught
graduate level courses in environmental toxicology. Dr. Jenkins has served on numerous panels
and task forces and has given testimony before Congress. He has been a consultant and member
of the Science Advisory Board of the U.S. Environmental Protection Agency. He has served the
National Research Council Panel on the Fate and Effects of Drilling Fluids, and task force to
develop revised Water Quality Standards for the State of Colorado, which he chaired.
Dr. Jenkins has expertise in contaminant fate and transport, contaminant bioavailability, and
contaminant metabolism and mechanisms of toxicity. He has experience with a wide range of
chemicals including metals, PCBs, dioxins, PAHs, pesticides including DDTs, and volatile
organic compounds. His experience includes the evaluation of toxicity caused by contaminated
sediments, sewage outfalls; urban runoff, drilling discharges, acid mine drainage, mine tailings,
groundwater discharges and the physical and chemical impacts of dredging activities. He has
worked extensively in large river systems in the Northeast, Midwest, Rocky Mountains and
British Columbia. He has also worked on coastal, estuarine, wetland and associated upland
systems through out the country.
-------
Peter F. Landrum, Ph.D.
Dr. Peter F. Landrum is a Senior Research Scientist at the Great Lakes Environmental Research
Laboratory of the National Oceanic and Atmospheric Administration in Ann Arbor, MI. He
received his B.S. in Chemistry from California State College San Bernardino and Ph.D. in
Pharmacology and Toxicology from the University of California, Davis.
Dr. Landrum spent the next two years working as a research associate for the University of Georgia
at the Savannah River Ecology Laboratory, Aiken SC. His research focused on the fate, transport
and bioaccumulation of polycyclic aromatic hydrocarbons in freshwater stream systems. He then
moved to the Great Lakes Environmental Research Laboratory as a research chemist to pursue
research in the bioavailability and bioaccumulation of organic contaminants by aquatic
invertebrates with and emphasis on the benthos.
Over the last twenty-one years, his research has examined the factors that affect the bioavailability
of organic contaminants including the role of dissolved organic matter in water borne exposures and
the influence of sediment characteristics on exposure to sediment-associated contaminants. More
recent work has examined the utility of whole body residue levels to determine the dose required
for contaminant toxicity. Dr. Landrum is recognized for his expertise on factors affecting the
bioavailability of contaminants and the toxicokinetics ofbenthic organisms.
Dr. Landrum was also active in teaching toxicology at Eastern Michigan University from 1984 -
1991. Dr. Landrum is active in the Society of Environmental Toxicology and Chemistry and served
on the board of directors from 1990 - 1993. Dr. Landrum is a fellow of the Cooperative Institute
for Limnology and Ecosystem Research (University of Michigan) and the Cooperative Institute for
Climate and Ocean Research (Woods Hole Oceanographic Institute) and has pursued cooperative
research with scientists from Clemson University, Michigan State University, University of
Michigan, Southern Illinois University, Wright State University, and Ohio State University. Dr.
Landrum received the SETAC Founders award in 1999, aNOAA Administrator's Award in 1999,
and was recently recognized by ISI as a Highly Cited Researcher in Ecology and Environmental
Science. Dr. Landrum has over 100 publications and served on the editorial boards for
Chemosphere and Critical Reviews in Environmental Science and Technology and is Associate
Editor for the Journal of Great Lakes Research.
-------
Lynn McCarty, Ph.D.
Dr. McCarty received B.Sc. and M.Sc. degrees from Brock University and a Ph.D. from the
University of Waterloo. He has spent over 25 years in examining various aspects of
environmental and human health, toxicology, and environmental contamination. This
includes employment at private consulting companies as well as the Health Studies Service of
the Ontario Ministry of Labour. Currently, he is the Principal of L.S. McCarty Scientific
Research & Consulting, an ecotoxicological consulting company and an adjunct professor in the
Department of Biology of the University of Waterloo.
He has been involved in a variety of projects examining human health effects and environmental
impacts in an assortment of situations and contaminants. This includes the preparation of over
two dozen human-health-based Ambient Air Quality Criteria documents for the Ontario
Government, production of seven ambient water quality guideline documents (both single and
groups of chemicals) for the Ontario and Canadian Governments, preparation and review of
computerized chemical dossiers for the CESARS database, and critical reviews of various
environmental quality guidelines, protocols, and risk assessments, including those associated with
the Canadian Environmental Protection Act. In addition Dr. McCarty has been an active team
member in a number of human-health and environmental risk modeling and assessment projects.
Dr. McCarty has been an invited expert at a number of workshops dealing with human and
environmental health sponsored by CNTC, SETAC, US EPA, and US Army Corps of Engineers.
He is a coauthor of two chapters in the second edition of the standard reference book
"Fundamentals of Aquatic Toxicology" (Rand, 1995). In addition to reports to clients,
he continues to publish in the primary scientific literature, contribute to book chapters (over 40
publications to date), make presentations at professional scientific meetings, and currently serves
on the editorial boards of Human and Ecological Risk Assessment and Journal of Aquatic
Ecosystem Stress and Recovery. The Ontario Government has recognized his scientific work by
awarding him the Ontario MOE Excellence in Research - Water Quality in 1990. His
toxicological interests are focused in the areas of Quantitative Structure-Activity Relationships
(QSAR), mixture toxicity, residue-based potency estimation, and environmental risk
management/assessment.
-------
Anne McElroy, Ph.D.
Dr. Anne McElroy is currently an associate professor at the Marine Sciences Research Center at
the State University of New York at Stony Brook where she teaches undergraduate courses on
oceanography, environmental problems and solutions, and environmental toxicology and public
health and graduate courses on aquatic toxicology and pollutant responses in marine organisms.
Her research focuses on how aquatic organisms interact with toxic chemicals in their
environment. Current projects are examining reproductive effects of sewage-derived
contaminants on fish and crustaceans, factors controlling bioaccumulation of organic
contaminants, factors influencing metabolism and genotoxicity of polycyclic aromatic
hydrocarbons, and development of new approaches to assessing the toxicity of contaminated
sediments. Dr. McElroy received a Science Bachelors in Aquatic Biology from Brown
University and her Ph.D. in Oceanography from the Massachusetts Institute of Technology-
Woods Hole Oceanographic Institution Joint Program. Before joining the faculty full time at
Stony Brook, Dr. McElroy served as Director of the New York State Sea Grant College Program
and was a faculty member in the Environmental Sciences Program at the University of
Massachusetts at Boston.
-------
James Meador, Ph.D.
Dr. James Meador is Team Leader for the Hazard Assessment and Risk Modeling Team (HARM) in the
Ecotoxicology Branch of the Environmental Conservation Division (NOAA Fisheries). His current
research interests include various aspects of the bioavailability, bioaccumulation, and toxicity of aquatic
contaminants. For many years he has studied the environmental factors that control the bioavailability of
contaminants and the amount that is bioaccumulated. He has been very interested in quantifying these
factors and using them to predict tissue concentrations in animals collected from the field. In conjunction
with this work, Dr. Meador has also examined the role of toxicokinetics in predicting bioaccumulation
and the differences that species exhibit in their response to toxicants. These studies have evolved to
include analysis of the toxicant response based on tissue concentrations. Dr. Meador and his group also
study habitat quality, dose-response relationships, and the effects of contaminants on salmonids and their
prey. A major function of the team is to generate and evaluate toxicity data to support the development of
water, tissue, and sediment quality guidelines that can be used to protect fish and invertebrate populations
from adverse effects. Before becoming the HARM team leader, Dr. Meador was the leader of the Toxic
Metals Research and Analysis Group. In 1989 Dr. Meador was a National Research Council Research
Associate at the National Marine Fisheries Laboratory in Seattle with Dr. U. Varanasi. Before persuing a
doctorate, he was a Staff Research Associate for 6 years at the Scripps Institution of Oceanography in San
Diego. Dr. Meador received a B.A. in Zoology from Humboldt State University, a M.S. in
Biology/Physiology from San Diego State University and his Ph.D. in Aquatic Toxicology from the
University of Washington. He has been a peer reviewer for several journals and organizations and
consults frequently with a wide variety of government agencies. Dr. Meador has authored or coauthored
35 peer-reviewed publications and several technical memoranda and reports. He has also given several
invited presentations on his work.
-------
Paul S. Price
Mr. Price is a director of The LifeLine Group, a nonprofit corporation developing software for the
assessment of exposure to pesticides and other substances. Mr. Price has more than 20 years of
experience in assessing exposure to chemicals for industry, government, and trade associations.
He has authored over 20 articles on exposure and risk assessment. Areas of interest include dose
reconstruction, aggregate and cumulative risk, integration of toxicity and exposure data using
simulation models, pesticide exposure, and exposures related to the consumption of fish. Mr.
Price has a Masters degree in Civil Engineering (University of Maryland, 1979) and a Bachelors
degree in Chemistry (University of Maryland, 1974).
-------
Harlee S. Strauss, Ph.D.
Dr. Harlee S. Strauss has a Ph.D. in molecular biology from the University of Wisconsin - Madison and
an A.B. in chemistry from Smith College. She was a postdoctoral fellow in biology at MIT and a
Congressional Science Fellow sponsored by the Biophysical Society. Dr. Strauss has more than 20 years
of experience in the areas of risk assessment and toxicology.
Dr. Strauss is currently the President of H. Strauss Associates, Inc., (HSAI) a consulting firm she founded
in 1988. She works on a broad range of projects, from site specific risk assessments to in-depth
evaluations of the toxicity of individual chemicals, to the development of frameworks for risk assessment
(e.g., for microorganisms). HSAI clients include private and public sector organizations and citizens
groups. Dr. Strauss has served as a member of the U.S. Army Science Board since 1994, and has
participated in studies regarding lead-based paint, groundwater and soil remediation at Army facilities,
Chem/Bio Weapons Defense, and the Range Rule (pertaining to unexploded ordinance). Dr. Strauss
served on the advisory committee for the Society for Risk Analysis Workshop "Key Issues in Carcinogen
Risk Assessment Guidelines" and on various peer review committees such as for the EPA Exposure
Factors Handbook and the Drake Incinerator risk assessment. She is a community member of the
Restoration Advisory Board of the U.S. Army's Soldiers System Center (Natick Labs) and an elected
Town Meeting member in Natick, Massachusetts.
-------
Richard J. Wenning
Mr. Richard Wenning is a Senior Manager in the San Francisco Bay Area (Emeryville),
California office of ENVIRON International Corporation. Mr. Wenning has over 15 years of
experience in applied health and ecological research using human health and ecological risk
assessment and risk-driven problem solving methods to evaluate and develop solutions to
environmental contamination. He is widely regarded as an expert in several aspects of chemical
risk assessment, including source identification, exposure modeling, food web modeling, and
probabilistic analysis. He has considerable experience in both human health and ecological risk
assessment of persistent, bioaccumulative organic chemicals such as the dioxins, PCBs, certain
chlorinated pesticides, and brominated flame retardants.
Mr. Wenning has served as an independent peer-reviewer of exposure and risk studies for
USEPA, state agencies, as well as environmental regulatory agencies in Australia and Italy. He
has provided expert testimony on exposure assessment and chemical fingerprinting for defendants
in product liability litigation and negotiations for allocation of cleanup costs at hazardous waste
sites. He has managed interdisciplinary teams of engineers and scientists on several multi-year
contaminated sediment investigations, HSWA chemical manufacturing facility assessments, and
contaminated property investigations. He has worked with clients throughout the U.S., as well as
in Australia, Canada, China, Italy, and New Zealand.
Mr. Wenning is active on the science advisory boards of several professional organizations and
international scientific conferences. He has published extensively in the scientific literature on
human health and ecological issues and assessment and remediation of contaminated sediments.
He is a senior editor of several scientific publications, including Contaminated Soil. Sediment &
Groundwater Magazine. Archives of Environmental Contamination & Toxicology. Ecotoxicoloev
& Environmental Safety, and Environmental Toxicology & Chemistry. In addition, Mr. Wenning
is a founding member and chairperson of the international steering committee for a SETAC-
sponsored Pellston Workshop on the use of sediment quality guidelines and related tools in
assessments of contaminated waterways (August, 2002).
-------
Appendix B
Peer Review Meeting: Overall Briefing
Summary and Attendance
-------
Peer Review Meeting:
Overall Briefing
Human Health & Shared Issues Charge Delivery
U.S. Environmental Protection Agency, Region 2 Offices
Room 27A, 290 Broadway, New York, NY
January 10-11,2002
MEETING SUMMARY
Thursday, January 10,2002 - Morning Session
Introductions
• Peer review panel introduced:
- G. Allen Burton, Wright State University
- Lynn McCarty, Independent Consultant
- Paul Price, Independent Consultant
- Richard Wenning, Environ Corp.
- Anne McElroy, SUNY, Stony Brook
- Harlee Strauss, Independent Consultant
- Kenneth Jenkins, Arcadis/JSA
- Peter Landrum, GLERL (absent due to family emergency)
- James Meador, NOAA
• RMW members in attendance stated their name and affiliation
Opening Remarks
Bill Muszynski (EPA Region 2) and Tom Creamer (USACE) welcomed the peer review
panel and provided perspective on the goals of EPA and USACE in convening the panel.
• Purpose of the peer review is to ensure that decisions regarding the HARS are based
on sound science
• Intent of the meeting was to provide historical context for the peer reviewers and to
allow all of the stakeholders an opportunity to present their views
• Goal is for the process to be transparent, therefore, all materials will be released to the
entire RMW
Overview of the Peer Review Process and Ground Rules
Nancy Bonnevie (Peer Review Coordinator, Battelle) provided a summary of the peer
review process as developed by the RMW and reviewed the rules established to maintain
the independence of the review. A copy of the presentation is available upon request.
• The peer review will be conducted in two phases, Human Health and Shared Issues
(Phase I) and Ecological Issues (Phase II)
• The peer reviewers will be asked to respond to each charge independently, however, a
primary goal will be for the panel to reach consensus on each of the issues identified
-------
HARS PEER REVIEW
Overall Briefing & Human Health & Shared Issues Charge Delivery
January 10-11,2002
• Throughout the course of the peer review process, the reviewers will not have contact
with any RMW member or with EPA or US ACE on any issue pertaining to the peer
review; Any questions etc. should be mediated through the Peer Review Coordinator.
Overview of the US ACE Dredging and Regulatory Program in the Region
Bryce Wisemiller (USACE) provided a summary of the New York District's Dredging
and Regulatory Program. Copies of this presentation are available upon request or at the
following web address: http://www.nan.usace.armv.mil
The Historic Area Remediation Site (HARS): Historic and Regulatory Background
Doug Pabst (EPA Region 2) provided a summary historic and regulatory background of
the HARS. Copies of this presentation are available upon request.
Overview of the Peer Review Binder and Related Materials
Nancy Bonnevie (Peer Review Coordinator, Battelle) walked through organization and
contents of the Peer Review Binder with the peer review panel. In addition, Tom
Creamer (USACE) reviewed the section containing the RMW white papers with the
panel to ensure that all panel members had copies of all white papers submitted.
RMW Presentations
RMW members who could not be present on Friday were invited to speak to the peer
reviewers. Speakers included:
• Dennis Suszkowski (Hudson River Foundation)
• Tom Wakeman (NY Port Authority)
Thursday. January 10.2002 - Afternoon Session
Closed discussion with Peer Reviewers Only (Facilitated by Battelle). Reviewed
charges, responded to questions and developed a schedule for proceeding on Phase I of
the review:
January 10-11,2002 Initial Peer Meeting/Phase I Charge Delivery (Panel &
RMW)
April 18-19,2002 Facilitated Consensus Meeting (Panel & Battelle, only)
May 20,2002 Draft Interim Consensus Report (Phase I) to EPA &
USACE
The schedule for Phase II will be reported as soon as it is finalized.
-------
HARS PEER REVIEW
Overall Briefing & Human Health & Shared Issues Charge Delivery
January 10-11,2002
Friday, January 11.2002
Synopsis of Proposed Approach for Assessing Risks to Human Health
Mark Reiss (EPA Region 2) provided a summary of the approach proposed for
addressing potential risks to human health.
USACE Recommendations and Preferences Regarding Risk Assessment at the
HARS
Todd Bridges (USACE -WES) provided a summary of the alternative approach
recommended by USACE. Copies of this presentation are available at the following web
address: http://www.nan.usace.armv.mil
Remediation Materials Workgroup (RMW) Presentations
RMW members were invited to present their comments and concerns to the peer
reviewers. Presenters included the following:
• Angela Cristini, Ramapo College (on behalf of Clean Ocean Action)
• James Tripp, Environmental Defense Fund
-------
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Appendix C
Scope of Work
-------
Scope of Work
Peer Review of the Proposed Bioaccumulation Testing Evaluation Framework
for Determining the Suitability of Dredged Material to be Placed at the
Historic Area Remediation Site (HARS).
1.0 Introduction and Project Background
Battelle has been contracted by the Environmental Protection Agency (EPA) Region 2 to
coordinate a scientific peer review of the Proposed Bioaccumulation Testing Evaluation
Framework for Determining the Suitability of Dredged Material to be Placed at the Historic Area
Remediation Site (HARS). The HARS was designated by the United States Environmental
Protection Agency (EPA) on September 29, 1997 simultaneous to the de-designation and
terminated use of the New York Bight Dredged Material Disposal Site [also known as the Mud
Dump Site (MDS)]. Prior to the HARS designation, EPA Region 2 committed to conduct a
public and scientific peer review process of its dredged material testing evaluation procedure
(NY/NJ HEP, 1996). Pursuant to that commitment, the EPA prepared a charge and presented
their testing evaluation framework to a scientific peer review panel. Based on the comments
received, EPA Region 2 has proposed changes to the testing evaluation framework, and is
seeking to have another peer review for the suggested modifications. The peer review will be
conducted in two phases, with the first phase focusing on technical issues related specifically to
the calculation of values protective of human health as well as those issues shared by both the
human health and ecological evaluations. The second phase will focus on those issues
specifically related to the calculation of values protection of ecological receptors. The peer
review will be coordinated in accordance with guidance provided in EPA's Peer Review
Handbook (EPA, 2000).
2.0 Task Summary
The primary task covered in this Scope of Work (SOW) is serving as a technical expert on the
peer review panel that will be convened to review the proposed changes to the HARS testing
evaluation framework. The level of effort is assumed to be approximately 152 hours based on
the proposed process as outlined below (Table 1; Figure 1).
Phase I: Human Health and Shared Issues
• All peer reviewers will gather for a two-day meeting in New York City on January 10 and
11,2002 for a briefing on the peer review process and the delivery of the Phase I charges,
specific to human health and shared issues. Representatives of EPA Region 2 (EPA), United
States Army Corps of Engineers (USACE), and the Remediation Workgroup (RMW) will be
present to provide background information and perspectives on the various technical issues
addressed by the charge.
• Following the initial briefing meeting each reviewer will have 60 days to independently
review the information provided and answer the specific charges relating to human health
and shared issues. At the end of this period, each reviewer will submit their independent
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responses to the charges to the Battelle Peer Review Leader. All submissions will be in
electronic (preferably MS Word) and hard copy format. As specified in Section 4.0 of this
SOW, it is required that the reviewers refrain from discussing the charges and their responses
with the other peer reviewers or any member of the RMW (including EPA and USACE)
during this independent review period. Specific technical questions and requests for
clarification that arise during this period should be directed to the Battelle Peer Review
Leader.
• All peer reviewers will reconvene for a two-day issue resolution meeting in New York City
to discuss the draft narrative report which will be compiled by Battelle based on the
comments received from the independent reviews. This document will be distributed to peer
reviewers at least one week prior to the meeting. During the meeting, a facilitated discussion
will be held to discuss outstanding issues and to resolve areas of conflicting opinion.
• Based on the outcome of the issue resolution meeting, Battelle will summarize the
recommendations of the panel regarding the human health and shared issues charge. This
information will be integrated with the results of Phase II to comprise a draft consensus
report for the overall peer review. This draft report will be presented to the panel for review
an approval before submission to EPA and the USACE.
Phase II: Ecological Issues
• All peer reviewers will gather for a one-day meeting in New York City for the delivery of the
Phase II charges, specific to the ecological issues. Representatives of EPA Region 2,
USACE, and the RMW will be present to provide background information and perspectives
on the various technical issues addressed by the charge.
• Following the initial briefing meeting each reviewer will have 60 days to independently
review the information provided and answer the specific charges relating to ecological issues.
At the end of this period, each reviewer will submit their independent responses to the
charges to the Battelle Peer Review Leader. All submissions will be in electronic (preferably
MS Word) and hard copy format. As specified in Section 4.0 of this SOW, it is required that
the reviewers refrain from discussing the charges and their responses with the other peer
reviewers or any member of the RMW during this independent review period. Specific
technical questions and requests for clarification that arise during this period should be
directed to the Battelle Peer Review Leader.
• All peer reviewers will reconvene for a two-day issue resolution meeting in New York City
to discuss the draft narrative report to be compiled by Battelle based on the comments
received from the independent reviews. This document will be distributed to peer reviewers
at least one week prior to the meeting. During the meeting, a facilitated discussion will be
held to discuss outstanding issues and to resolve areas of conflicting opinion.
• Based on the outcome of the issue resolution meeting, Battelle will summarize the
recommendations of the panel regarding the ecological issues charge in an interim report.
-------
This information will be integrated with the results of Phase II to comprise a draft consensus
report for the overall peer review.
Final Report and Presentation of Results
• The reviewers will be asked to review and approve the draft consensus report. This report
will be presented by the Battelle Peer Review Leader and the peer review panel at an RMW
meeting. At that meeting, panel members will be asked to respond to any questions from the
RMW regarding the consensus report and the conclusions reached.
3.0 Deliverables and Schedule
Deliverables required for this SOW include the following:
• Written response to the charges presented. Where necessary, background and reference
material supporting the response should be provided as well. Each reviewer will be expected
to respond to all questions posed in the charge to the best of their ability. If a reviewer feels
unqualified to respond to a particular question, the Battelle Peer Review Leader should be
informed as soon as possible.
• Participation in all meetings identified in Section 2.0. Whenever possible, attempts will be
made to overlap meetings required for Phase I and II to reduce the travel time required.
• Written review comments on the interim reports for each phase as well as the final combined
consensus report.
The initial meeting for Phase I of the peer review will take place on January 10 and 11,2002 in
New York City. Specific dates for subsequent meetings and deliverables will be established at
that time, however, anticipated time frames for each step of the peer review process are outlined
in the flow-chart in Figure 1.
4.0 Reporting and Communications
Peer reviewers will report to the Battelle Peer Review Leader:
Nancy L. Bonnevie
Battelle
397 Washington Street
Duxbury, MA 02332
781-952-5384
As previously noted, with the exception of the meetings specified in Section 2.0 of this SOW, it
is required that the reviewers refrain from discussing the charges and their individual responses
with the other peer reviewers or any member of the RMW. All specific technical questions and
requests for clarification that arise should be directed to the Battelle Peer Review Leader.
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Table 1. Level of Effort for Peer Review Panel
Peer Review Task
Estimated
Level of Effort
Basis
Phase I: Human Health & Shared Issues
Overall Briefing and Human
Health & Shared Issues Charge
Delivery
16
2-day meeting per RMW
Independent Scientific Review of
Human Health & Shared Issues
40
Assumption
Issue Resolution of Human
Health & Shared Issues
16
2-day meeting per RMW
Human Health & Shared Issues
Consensus Report
8
Assumption
Phase II: Ecological Issues
Ecological Issues Charge Delivery
8
Assumption
Independent Scientific Review of
Ecological Issues
24
Assumption
Issue Resolution of Ecological
Issues
16
2-day meeting per RMW
Ecological Issues Consensus
Report
8
Assumption
Final Consensus Report
Review
8
Assumption
Presentation to RMW
8
1-day meeting per RMW
Total Level of Effort
152
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llBaffeile
. .. Putting Technology To Work
Date June 5, 2002
To Peer Reviewer
From Nancy Bonnevie
Battelle
397 Washington Street
Duxbury, MA 02332
Subject Conflict of Interest Inquiry
You have been selected to serve on the peer review panel for the review of the Proposed Bioaccumulation
Testing Evaluation Framework for Determining the Suitability of Dredged Material to be placed at the
Historic Area Remediation Site (HARS). Your participation in this review will be greatly appreciated.
Prior to initiating the process we would like to be aware of any personal affiliations or involvement in
particular activities that could pose a conflict of interest or create the appearance that you lack impartiality
in your involvement for this peer review. Although your involvement in such activities is not necessarily
grounds for exclusion from the peer review it should be brought to my attention so that we can discuss
any potential implications. Affiliations or activities that could potentially lead to conflicts of interest are
listed below.
• Work or arrangements concerning future work in support of industries or other parties that could
potentially be affected by regulatory developments or other actions based on material presented in
the document (or review material). Examples of such industries or other parties include state or
federal regulatory agencies involved with dredging issues in New York and New Jersey Harbor or
environmental or shipping organizations with perceived interests in the result of the project.
• Personal benefit (or benefit of your employer, spouse or dependent child) from the developments
or other actions based on the document (or review materials) you have been asked to review.
• Membership or inferred involvement in any organization that possesses extreme points of view
on environmental or shipping interests.
• Bias or extreme points of view towards either environmental or shipping interests.
• Bias or extreme points of view related to the subject matter of scientific material being reviewed.
• Previous involvement with the development of the document (or review materials).
• Financial interest held by you (or your employer, spouse or dependent child) that could be
affected by your participation in this matter.
• Financial relationship you have or have had with EPA (such as research grants or cooperative
agreements) related to dredging issues.
Please contact me at (781) 952-5384 to discuss any potential conflict of interest issues as soon as
possible. Please sign and return this no later than December 21,2001.
Name
Date
-------
Figure 1. Proposed Peer Review Process
PHASE I
HI'MAN HEALTH
RMW Meeting:
Shared Issues
(2D
luman Health &
Charge Input
1
f 60 Days
Peer Reviewer Meeting: Overall Briefing 1
and Human Health & Shared Issues
Charge Delivery (2 Days)
Independent Scientific Review of Human!
Health & Shared Issues
60 Days
Peer Review Meeting: Issue Resolution
for Human Health & Shared Issues
Human Health & Shared Issues
Interim Report
PHASE II
ECOLOGICAL EFFECTS
RMW Meetings:
Ecological TEF Development
X 30 Days
Prepare Ecological TEF
for Example Compounds
120 Days
RMW Meeting: Ecological Issues
Charge Input
•w
Peer Reviewer Meeting: Ecological
Issues Charge Delivery
i
Independent Scientific Review of
Ecological Issues
60 Days
Peer Reviewer Meeting: Issue
Resolution for Ecological Issues
(2 Days)
Interim Report: Ecological Issues
(30 Days)
Human and Ecological
Consensus Report
RMW
Briefing
—r
Prepare Draft TEF
(3 Months)
t
Review/Finalize Responsiveness
Report
(2 Months)
RMW Review and Discussions
Q
RMW Only
Q
Peer Reviewers Only
Q
RMW & Peer
Reviewers
Revise Draft TEF
(21 days)
Public Comment
(60 Days)
I
Finalize Technical Resolution and
TEF Amendment*
Finalize TEF
-------
>
13
73
rt
S.
Appendix D
Premeeting Comments
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Comments Received from
G. Allen Burton, Jr., Ph.D.
March 16,2002
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HARS Review: Human Health Risk Framework
G. Allen Burton, Jr., Ph.D.
March 16,2002 Draft
Reviewer Note:
My expertise is in the assessment of aquatic ecosystem impairment, particularly where
sediments and stormwaters dominate as stressors. Many of the human health risk
questions and issues of concern in this first phase of HARS review will also be an issue
(and likely of greater consequence) in the subsequent ecological risk review. Those
similar issues will be dealt with in greater detail and more extensive peer-reviewed
literature documentation can be provided.
Reviewer responses to each HARS Scientific Peer Review Questions (italics) are
provided below. All comments are in the context of the HARS Scientific Peer Review
Charges and Questions, unless otherwise noted. Additional comments are also provided
where appropriate.
Overall Process
1. Throughout the proposed process, there are various uncertainties introduced.
Please identify the key areas of uncertainty that need to be addressed. Are there
additional data sources or parameters that could be used to address these areas?
What methods are available for describing and accounting for these uncertainties
in the calculation ofHARS-Specific Values? Of the methods available, which
would you recommendfor consideration and why? Please consider the
implications of implementing these methods in the regulatory framework. Please
include an evaluation of probabilistic and deterministic methods in your
discussion.
A common line of contention throughout the HARS process (and in most other risk
assessments) is that of uncertainty associated with whether or not specific chemicals are
causing significant adverse biological effects. The challenge continues to be that of
building a credible scientific foundation to support the traditional regulatory framework
that is driven by chemical specific values. The logic of both developing and using this
simplistic system is apparent at first glance, and has been adopted by regulatory agencies
throughout the world. However, it is now well established via the peer-reviewed
literature and billions of dollars in litigation, that scientific uncertainties associated with
chemical specific effect values make their application a contentious challenge, at best. It
is not feasible at this point, for purposes of this initial review, to adequately document all
the reasons why chemical specific risk values often lead to significant uncertainties;
particularly in dealing with sediments and when in complex, dynamic ecosystems where
multiple stressors exist (for example (in regard to ecosystem assessments) see Burton
1991,1992; Burton and Pitt 2001; Burton et al, 1996,2000,2001,2002). In simplistic
terms, the significant uncertainties can be grouped and described as follows:
1. Threshold Value Derivation: For Human Health, these values originate
primarily from rodent effects data where safe levels are modeled and
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2
extrapolated to humans. For aquatic species, these values originate from the
water quality criteria. Water quality criteria are derived from laboratory
toxicity tests, which produce species-specific thresholds at one life stage,
under constant exposure conditions, in chemical spiked, solids-free waters (of
different chemistry from the test site). For either humans or aquatic species,
the chemistry of the target compound, its bioavailability, the exposure
dynamics, co-occurrence of other stressors, and the test species differ from the
receptor of concern at the test site.
2. Bioavailability: There has been reasonable success in predicting the
bioavailability (toxicity) of some divalent cationic metals in freshwater
systems using hardness to normalize. Recently there has been success with
metals using the Biotic Ligand Model. In sediments, our ability to predict
bioavailability of metals or organics is not as effective, and a point of strong
debate as evidence by the upcoming Pellston Workshop on sediment quality
guidelines and opposition of many to the U.S. EPA's equilibrium partitioning
approach. Again, the regulatory approach is to predict bioavailability on a
chemical-specific basis, while organisms are affected by mixtures of
chemicals. This shortcoming is addressed to some degree through the
occasional requirement for laboratory toxicity testing.
3. Mixtures: Individuals are affected by mixtures. Discerning if adverse effects
on a receptor is due to one chemical, when it is exposed to many stressors is
rarely possible, hence causality is often a significant uncertainty. In addition,
when chemicals are mixed together, their resulting biological effects can be
modified by interactions. These interactions can be characterized as being
additive, antagonistic, or synergistic in nature. This straight-forward reality is
indeed a critical one when predicting risk using chemical-specific values for
humans or ecological receptors. The current regulatory process assumes
additivity, which allows the chemical-specific assessment approach to work.
However, there is a growing body of literature showing many exceptions to
the additivity assumption (e.g., see citations in Burton et al. 2000). This has
huge implications on the risk prediction! Given the complexity of
experimental design for mixture studies, this will be an area of great
uncertainty for many, many years to come. Other mixture uncertainties
include the fact that organisms are not exposed to each chemical of a mixture
equally, some chemicals will be more bioavailable than others and also
available for uptake differing periods of time due to their own unique
physicochemical characteristics and sources. Resulting interactions vary with
chemical type, concentration, temperature and target species. The incredible
complexity of trying to predict true risk based on chemical specific values can
be envisioned by this simplistic example:
During a 24 hr period a fish swimming through the New York harbor may
be exposed to Chemical A for 10 minutes at a concentration exceeding a
threshold value when it is 100% available, then for 10 hours exceeding
threshold values (but only 10% available), and the remaining day at half
the threshold value (availability ranging from 0 to 100%).
Simultaneously, the fish exposures to Chemical B are the opposite, and
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3
Chemical C, D, etc. each have their own unique exposure patterns. In
addition, all the threshold values used above were calculated based on 2
day or 7 day constant exposure... When the fish is exposed to a mixture of
Chemical A + B at low concentrations, and at low temperatures there is a
synergistic effect (greater than expected toxicity); however at higher
concentrations, there is an antagonistic effect. This temperature and
concentration specific interaction pattern is reversed with Chemical C + D.
This seemingly odd interaction pattern has been shown to exist with multiple
metals in fish and likely occurs with organics. In summary, it is apparent it is
not feasible for an accurate prediction of mixture effects, for one or multiple
exposure scenarios, using a chemical specific approach.
4. Exposure vs. Effects: From the above discussion of three uncertainty issues it
should be apparent why it is difficult to accurately predict exposure and
effects in dynamic ecosystems, possessing multiple potential stressors.
However, the science is improving in this regard, as we are now able to better
characterize movements of benthic and pelagic organisms, thus can better
characterize their potential exposure to areas of contamination. The
challenge, however, is to relate a significant adverse effect to a specific
exposure period, based on laboratory derived, chemical specific threshold
effects. As stated above these values are derived for one (perhaps 2) exposure
periods at constant concentrations. There is a growing body of literature that
is showing pulse exposures to single chemicals or mixtures may result to
either increased or decreased toxicity. As with the above mixture example,
this makes prediction of field effects very tenuous because the underlying
assumptions have been proven wrong or relevant, exposure patterns simply do
not apply. In addition, organisms are exposed to chemicals via multiple
compartments (i.e., surface water dissolved, surface water
particulate/colloidal, sediment, interstitial water, upwellings of groundwater,
and food). Each exposure compartment has potential chemical stressors at
differing concentrations, with differing bioavailabilities and the organism is
exposed to each compartment for varying time periods.
Given these confounding factors, the physical, chemical and ecological complexity of the
HARS site, and the importance of this risk assessment process, it is difficult to justify
reliance on the currently proposed, chemical specific value approach. Rather, a tiered
assessment approach should be used that removes these huge scientific uncertainties by
using a more straight-forward, biologically-based approach (Burton 2001). If actual,
adverse effects are observed in Tier 1 screening studies, then the magnitude (system
significance) of the adverse effect can be quantified in Tier 2. If deemed necessary, the
dominant stressors can be identified (causality) in Tier 3, requiring more chemical
fractionation with biologically-based effects testing (e.g., Toxicity Identification
Evaluation). However, given the flaws of the chemical-specific approach (described
above) and potential creation of artifacts from sample handling, manipulation and
laboratory testing (e.g., ASTM 1994; USEPA 2001; Burton 1992), it is difficult to
envision how Tier 3 will help the HARS management process of ecological effects.
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4
The challenge of determining human risk from consumption of contaminated seafood is
large. The questions outlined for the HARS Scientific Peer Review are good ones that
address the challenge. Many of the issues addressed above focus on determining
ecological risk, but are directly and indirectly affecting the human risk process. In this
regard, the pre-defined HARS Scientific Peer Review process seems flawed. It would
have been more logical from a scientific perspective, to address the many questions
(below) pertaining to site chemistry and organism exposure (i.e., ecological risk) first,
then discuss extrapolation of tissue values to human risk predictions. The reality is that
human risk predictions will be relatively crude, and be largely based on significant
assumptions of tissue concentrations of a suite of chemicals (on a chemical specific
basis), their adverse effects, and ingestion rates. The HARS risk management process
will be most efficient and effective, if the human and ecological risk assessment
uncertainties and design issues are considered together, in a coordinated approach.
Because we can assess adverse effects to key receptors directly in aquatic ecosystems (as
opposed to extrapolating chemical specific values for human risk), the ecological risk
assessment process can be done with fewer assumptions and uncertainties, and in a
straight-forward, tiered manner (above). This process should be done to assess ecological
risk accurately. These risk predictions should be compared to the human health risk
assessment, comparing shellfish and fish tissue concentrations between the two risk
assessments. In the ERA process key receptors (and assessment/measurement endpoints)
will be identified, along with associated toxicity thresholds. The tissue concentrations
that occur in these receptors at toxicity thresholds should be identified, as labeled as co-
occurring threshold tissue concentrations (causality has not been established, so the tissue
concentrations cannot be used as a risk driver). In the Human Health Risk Assessment
process, the receptors of concern have been identified (principal shellfish and fish species
that are ingested). However, following the ERA process it may be documented that some
species of concern to human health, are not primarily exposed at the HARS, thus should
not be in the HARS Human Risk Assessment process (HRAP). This exposure
characterization, of course, could also occur in the HRAP. Either way, it must be done.
If the ERA finds adverse effects are occurring and the co-occurring tissue concentrations
are below predicted Human Risk thresholds, then the ERA toxicity thresholds will drive
management of HARS. Conversely, if the Human Risk thresholds are below the ERA
co-occurring tissue concentrations, then these chemical-specific tissue concentrations
may drive the management of HARS.
Details on the tiered, biologically-based, toxicity threshold approach for the ERA will be
described during the HARS ERA Scientific Peer Review process. It is alluded to here,
because of the significant overlap of issues and the regulatory process being used. Many
of the issues being questioned below, become irrelevant, if a biological effect approach is
used rather than a chemical specific value approach. Simply put, if significant toxicity or
bioaccumulation is observed then the test material is unacceptable (as in current U.S.
Army Corps of Engineers (USACE)/U.S. Environmental Protection Agency (USEPA)
guidelines for dredged material testing). However, given the uncertainties associated
with laboratory testing (described above, e.g., constant exposures, sampling/manipulation
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5
affects on bioavailability), field validations must occur that measure toxicity and
bioaccumulation at the test site. Exposure-Effect linkages can be made in the field using
a combination of caged animal exposures, passive uptake devices (e.g., semi-permeable
membrane devices), and by direct sampling of indigenous species. Both options should
be tied to target receptor movement patterns, which must be defined for both the ERA
and HRAP. The biologically based approach removes the need to determining
causality, identifying the bioavailable fraction, and chemical specific threshold levels.
However, chemical specific values (based on the mammalian toxicity literature values
and FDA guidelines) will still need to be used for the HRAP.
Proposed Additions to Analvte List: Alkylated PAHs
2. Is measurement of the 16 priority pollutant PAHs (i.e. parent PAHs) sufficient for
characterizing the risks associated with the total PAH bioaccumulated by
organisms exposed to dredged material proposedfor placement at the HARS?
Does measurement of the alkylated compounds significantly improve risk
assessment of PAHs?
Initially, it is not adequate to simply assess only 16 parent PAH compounds to determine
the risks associated with total PAHs bioaccumulated by organisms. While there is much
less toxicological information available for alkylated compounds, there is adequate
information to know that some of these compounds pose a risk. For example, in recent
years Dr. Peter Hodson and others (University of Quebec) have been studying the effects
of the many compounds associated with pulp and paper mill effluents, including retene.
This compound is fairly well characterized. The measurement of the alkylated
compounds significantly improves the validity of the risk assessment, since it is likely
these compounds are present. Whether or not alkylated compounds can be assessed with
adequate power and sensitivity and whether or not they pose a significant risk cannot be
determined in a credible manner at this time. Therefore, it is important that initial phases
of the HARS assessment process include broad scale assessments of all PAHs and their
related compounds. The distribution of PAHs will be very source and age dependent and
of course are metabolized to varying degrees by different species.
3. Is the proposed adaptation of EPA Method 8270 (Appendix D) acceptable and
appropriate for regulatory decision-making? If not, what is an acceptable and
appropriate method?
Will provide response later, following additional research.
4. Under what specific conditions would the testing for alkylated PAHs for a
particular project be appropriate and warranted?
They likely always occur in urban watersheds, thus are present at HARS and a potential
stressor to receptors. However, the standard suite of 16 parent compounds likely serve as
an adequate surrogate for their occurrence in most circumstances, negating the need for
analyses. IF reliable toxicological information is available for specific alkylated
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6
compounds, then a screening survey should be conducted for the presence of those
compounds and determination whether it should be included in the HRAP.
5. What uncertainties would be introduced within the analysis of risk should
alkylated PAHs be included? What steps could be taken to account for these
uncertainties in decision-making? Given the likelihood the method for using non-
detects (as described in EPAJCENAN, 1997) will result in an overestimate of risk,
what are the implications?
As discussed above, if alkylated PAH presence is confirmed in sediments and in tissues
of receptors collected from HARS, and if reliable toxicological information is available,
then they should be included. Inclusion of these compounds into the HRAP or ERA
without meeting these criteria will produce unwarranted uncertainties. So, no steps
should be taken into account if the above criteria are met. They should not be included
in the HRAP or ERA without documentation of presence, nor should any uncertainty
factors be incorporated to make up for this deficiency. The non-detect issue has been
well addressed in the literature, as cited by others on this Peer Review Panel. State-of-
the-science consensus on non-detects should be incorporated into HARS assessment
process.
Proposed Additions to Analvte List: Organotins
6. It is recognized that additional methods have been used for the analysis of
organotins (e.g., Krone et al., 1989). Will the proposed analytical method (Rice
et al., 1987) provide adequate data of sufficient quality to assess relevant risks
from organotins? If not, please provide recommendations.
Yes, these methods do provide adequate data for incorporation into the risk assessment
process. Note comments by Panelist Jim Meador who is a recognized expert on
organotin analyses and effects.
7. What special QA/QCprocedures should be implemented to ensure the quality and
usability of the organotin data?
Sediment sampling, storage, manipulation and analyses can have significant impacts on
chemical concentrations, speciation and bioavailability (USEPA 2001). New guidance
available from the USEPA (2001) should be closely followed and documented to ensure
proper QA/QC procedures are met. For guidance unique to organotins, see comments by
Jim Meador.
8. Under what specific conditions would the testing for organotins for a particular
project be appropriate and warranted?
A study by Moore et al. 1991 stated a critical tissue concentration of 6.3 ug/g. If
concentrations near this level are measured in tissues of benthic species residing at
HARS, then organotins should be further assessed.
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7
Provosed Additions to Analvte List: Coplanar PCB Congeners
9. If the approach for evaluating dioxin is modified, should it include the
contribution ofPCBs with dioxin-like activity as proposed? If so, how?
PCBs with dioxin-like activity should be included in the HRAP. This can be done
following the WHO 1998 approach, which states specific PCBs that possess this activity
(10 congeners that are tetra-hepta substituted, including PCB 77, 81,105,114,118,123,
126,156,157,167,169, and 189). Their potencies are added into the total dioxin load in
the tissues and compared to toxicity equivalents. Total PCB classification can be
described following NOAA 1989 (see EPA/CENAN, 1998).
Comparison to Reference
10. Please consider the policy for assigning values (at one half the detection limit) to
tissue residues that are reported as "
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8
necessary to do so but this uncertainty adds to the argument made above for a field-based,
biological approach where the predictions, based on these assumptions can be verified.
The theoretical approach described above for Kow may be acceptable for low to
moderate log Kow values (i.e., < 6); however the uptake-Kow relationship breaks down
at higher values.
13. Given the increased hydrophobicity of alkylated PAHs, is the use of the correction
factor associated with the corresponding parent an appropriate approach for
estimating steady state residues of alkylated PAHs? If not, please elaborate.
See earlier comments regarding validity of using alkylated PAHs. If those criteria met
and if log Kow values below 6, then a correction factor may be appropriate.
14. For the DDT derivatives and dieldrin, please comment on the appropriateness of
using M. nasuta data rather than N. virens-specific data in the estimation of
steady state multipliers.
Perhaps, if feeding and bioaccumulation patterns are the same. This should be
documented prior to cross use, however. It should be recognized, however that if they
are feeding from differing sediment depths, their exposure will be different. This and
their predator-prey relationship with other receptors must be considered in the food web
contamination potential (risk assessment process).
15. Are the approaches taken to adjust organic contaminant bioaccumulation data to
steady state adequate? Do the proposed multipliers agree with previously
published studies (i.e., do they appear reasonable)? If not, please elaborate.
See above comments regarding exposure vs. effects dynamics and uncertainties and the
lack of steady state. Further field validation of steady state assumptions is essential.
16. What are the major sources of uncertainty associated with the approaches? What
alternative approaches would reduce the uncertainties? How could these
uncertainties be described and accounted for in decision-making?
See above comments regarding exposure vs. effects dynamics and the numerous
uncertainties of the chemical specific threshold value approach. Other uncertainties
include extrapolation of literature based effect thresholds to other species, differences in
feeding (thus contaminant exposure), metabolism of chemicals of concern, adaptation,
acclimation, and an understanding of which life stage is the most critical. Finally,
chemical weathering and sequestration has been shown to have significant impacts on
bioavailability, thus uptake and toxicity. Chemicals at HARS vary from in age from 1
day to many years, thus predicting availability is extremely tenuous.
Given these many uncertainties and as discussed under no. 1 above, there is a need for a
more straight-forward, field biology based approach with less associated uncertainty. A
three component assessment approach would greatly reduce uncertainties. The
integrated components of this process are: 1) Establish a validated spatial exposure
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9
models for key receptors; 2) Validate model with targeted analyses of indigenous biota
presence and tissue concentrations (and SPMD concentrations), matched with water and
sediment concentrations or target chemicals; and 3) Verify biological exposure and
effects using site specific measures of benthic invertebrate community structure and
toxicity and bioaccumulation in caged organisms (sediment vs. water column
compartments). Then, accurate tissue numbers can be used for deriving HARS specific
values for the HRAP and the remainder of the data can be used in the ERA process.
Adjustment to Steady State: Metals
17. In your opinion, is the methodology followed to derive the steady state multiplier
for non-essential metals (i.e., a factor of three) scientifically appropriate
(Appendix G)? Please elaborate. Do you have any recommendations of
additional or alternate methodologies or information that can be used to either
supplement or replace the proposed method?
Will provide response later, following additional research.
Human Health Evaluations: Overall
18. Please comment on each factor listed above (and in Table 5) as to its
appropriateness for use in the equations listed above. Would you recommend
additional factors? Would you change or modify the equations as written above?
If so, how?
The validity of the assumption "that suspended and dissolved constituents of dredged
material do not persist in the water column following release from the barge" should be
verified. It is difficult to imagine that a significant loading of clays that do not readily
settle, colloids and dissolved contaminants do not persist in the water column and
potential affect planktonic and pelagic species via sorption or ingestion. To properly
assess this phenomenon is no easy task, thus the assumption is questionable.
For cancer potency factors and reference doses of alkylated PAHs, use of parent
compound values should be done with caution. See above comments on verification of
alkylated data. Note that PAHs are largely metabolized and excreted by fish, and not
sequestered in fillets. While they may pose an ecological risk, to organisms eating whole
fish where there is recent exposure, the possibility of human uptake appears questionable.
In regards to seafood consumption values, the validity of the NJMSC study to HARS
should be reviewed. This is a critical number that must be matched with an accurate
assessment of catch exposure to HARS. See discussion of causality above.
While the USEPA's policy is to use a daily consumption rate for 70 years, we all know
that is excessive and grossly unrealistic. This mandated conservative approach adds
uncertainty that should be challenged, rather than routinely accepted simply because it is
policy. It is simply not reality.
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As discussed above, an accurate exposure characterization is essential. It represents one
half of the risk assessment process. While the derivation process used to achieve a factor
may be mathematically correct based on the stated assumptions, it is overly simplistic and
cannot represent reality. A more accurate site use factor must be developed using the
approach outlined above (e.g., response no. 16). Site use will vary between species
(benthic and pelagic) and on a daily, seasonal, and annual basis.
The whole body to fillet factor questionable based on the following realities: 1) Some
target subgroups (e.g., Asian, Afro-American) may consume whole fish; 2) lipid
concentrations in fish vary within (age, sex, season) and between species.
Food web models that employ trophic tranfer factors can vary orders of magnitude in
their resulting predictions, and can vary orders of magnitude from actual site specific
findings. This is due to a host of factors, many of which have been discussed above. In
addition, some invertebrates can metabolize PAHS, in addition to the metabolism by fish,
making their trophic transfer prediction more difficult. Interpretation of metals tissue
concentrations to assess bioaccumulation/bioconcentration factors or to predict food
chain effects are particularly problematic and variable. Any trophic transfer predictions
must be field validated as described above.
Will provide additional response on metal issues, following additional research.
19. Are the methods used to derive the human health exposure parameters and
assigned values discussed in Section E appropriate (please review the referenced
appendices)? If not, please elaborate. How should these factors be factored into
the risk analyses and decision-making?
See comments in no. 18.
20. Is the approach taken to relate fish whole body and fillet concentrations
scientifically appropriate? If not, what method would you recommend?
See comments in no. 18.
21. Could the analysis be improved by focusing on key fish (seafood) species at the
HARS? What characteristics should be used to select these key species?
Yes. Again the HRAP and ERA process should be integrated here, as both need this
information; however the decisions for selecting "key" fish may vary with each. Key
selection characteristics include: magnitude of consumption, duration/frequency of
HARS exposure, and importance to food web (ecologically and as a contaminant transfer
agent).
22. In your opinion, is the approach for assuming total metal to be in the most toxic
form appropriate and reasonable? Should metal speciation/complexation be
considered in the assessment of metals bioaccumulation, trophic transfer, and
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11
human health risks? Is the proposed approach for evaluating methyl mercury
appropriate? Are there alternative analytical or risk assessment techniques
available that would improve the risk assessment of metals? Is the multiplier
proposed for adjusting measured concentrations of arsenic appropriate and
reasonable?
It is well recognized that total metal concentrations have little to no relationship to
bioavailability. Metal speciation/complexation must be considered in the assessment of
bioaccumulation, trophic transfer and human health risks. Again, this is a case where
coordination of the HRAP and ERA process is needed, as many organisms accumulate
metal concentrations that can be quite high, without adverse effects, depending on
numerous factors. Metal tissue levels are extremely difficult to interpret in terms of
adverse responses. As discussed above, the relatively recent Biotic Ligand Model shows
a high degree of promise for assessing ecological effects. The acceptability of the arsenic
multiplier will require further review.
23. Is the assumption that the potency of alkylated PAHs can be estimated by the
potency of the parent PAH appropriate? Is this assumption likely to result in an
under- or overestimate of the risk associated with the alkylated PAHs?
This is a major uncertainty and should only be done on compounds where the relationship
has been documented, as discussed above.
24. Please comment on the potential for human exposure to PAHs through
consumption of finfish and other seafood.
See above discussion, no. 18.
25. What are the major sources of uncertainty associated with the approaches
described in Section E? What alternative approaches would reduce the
uncertainties? How could these uncertainties be described and accounted for in
decision-making?
See above discussion.
26. What is your recommendation for evaluating the potential toxicity of organotins?
Should they be evaluated as individual compounds? Summed as total? Should
there be some consideration of relative toxicity?
The focus should be on tributyltin, as that is the compound we know the most about.
27. Please comment on the appropriateness of the proposed approach for converting
and using the analytical data for alkylated and parent PAHs to estimate risk from
all PAHs.
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Human Health Evaluations: Comparison to HARS-Specific Values
28. Do you believe that the "disaggregate " modeling discussed above (and shown in
Figure 4) for estimating human health HARS-Specific Values for lead is
appropriate? Would you recommend an alternative risk assessment method be
used given the information and data available? Do you believe the method
described has appropriately taken uncertainty into account? Please elaborate.
Additional response will be provided following further research.
Human Health Evaluations: Consideration of Combined effects
29. In your opinion, are the methodologies and equations described above
appropriate for estimating total carcinogenicity and combined non-cancer
impacts of contaminant mixtures accumulatedfrom dredged materials proposed
for use as Remediation Material at the HARS?
Yes, but I have limited expertise in this area.
30. Is the conceptual model for evaluating fish exposure to dredged material at the
HARS and human exposure through ingestion of seafood appropriate and
reasonable? How can the uncertainties associated with the assumptions in this
conceptual model be reduced? Please consider the spatial and temporal elements
of exposure in your discussion.
No, the model is not appropriate or reasonable, based on the above discussions. See
above discussions for discussions on how to reduce uncertainties and accurately
characterize exposure.
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Comments Received from
Kenneth D. Jenkins, Ph.D.
March 29,2002
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Proposed Bioaccumulation Testing Evaluation Framework for Assessing the
Suitability of Dredged Materials to Be Placed at the Historic area Remediation
Site (HARS)
Response to Charges to the Peer Review Panel
Kenneth D. Jenkins, Ph.D.
March 29,2002
1. Key areas of uncertainty and the use of probabilistic risk assessment:
My primary concern with the overall framework is that each of the parameters used in estimating
potential for exposure of humans to HARS related chemicals are based on conservative point
estimates. These point estimates do not adequately account for the uncertainty and variability
associated with each parameter. Moreover, the methods presented in the framework document do
not explicitly address these factors or evaluate how they may affect exposure estimates. Instead,
uncertainty and variability are addressed indirectly by incorporating multiple conservative
assumptions in the development of each of the point estimates.
The effects of uncertainty and variability on exposure estimates can be addressed more explicitly
by incorporating a probabilistic analysis such as that proposed by the USACE. This approach
would provide a more rigorous basis for explicitly describing the uncertainty and variability
associated with the various exposure parameters. A probabilistic approach would also allow the
evaluation of their importance relative importance when estimating the potential for exposure of
humans to HARS related chemicals.
Proposed Addition to the analvte list: Alkylated PAHs
2. Is measurement of the 16 priority pollutant PAHs sufficient for characterizing risk
associated with total PAHs?
Measurement of only the 16 priority pollutants may substantially underestimate the mass of PAHs
accumulated in biota that have been exposed to sediments with petroleum hydrocarbon
contamination (Irwin et al., 1997). This could result in a significant underestimation of exposure
to PAHs. Inclusion of the alkylated homologs, which usually account for a much higher
proportion of the PAH mass than the parent compounds, would address this underestimation and
thus provide a more accurate estimate of the exposure to total PAHs. However, given uncertainties
associated with estimating steady state or trophic transfer for alkylated PAHs, and the lack of
toxicity data, it is unclear if their inclusion in the exposure estimates will actually reduce the
uncertainty in estimates of risk to human health. (See comments on Charges 5,13,23 and 24)
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3. Use of EPA method 8270:
I have had limited experience with the analysis of PAHs and I do not feel qualified to address this
issue.
4. Circumstances where testing for alkylated PAHs may be warranted:
If alkylated PAHs are to be incorporated in the analysis, they should be included in all projects
where parent PAHs are analyzed. I have commented on the appropriateness of incorporating
alkylated PAHs in the analysis in my response to Charge 23.
5. Uncertainties associated with the inclusion of alkylated PAHs a risk analysis.
As the bulk of PAHs may often exist in alkylated form (Irwin et al., 1997), including alkylated
compounds in the risk assessment would increase the mass of compounds analyzed and more
accurately reflect exposure conditions. However, including alkylated PAHs would increase
uncertainty in risk analysis, since bioaccumulation and persistence, metabolism, toxicity,
carcinogenicity, and sublethal effects of alkylated compounds are often unknown (Baussant, et al.,
2001; Irwin et al., 1997). (See comments on Charges 2,13,23 and 24)
Proposed Addition to the Analvte List
6. Method for organotins analysis:
I have had no experience with the analysis of organotins and I do not feel qualified to address this
issue.
7. QA/QC procedure for organotins:
I have had no experience with the analysis of organotins and I do not feel qualified to address this
issue.
8. Circumstances where testing for organotins may be warranted:
Data summarized in Appendix E, although limited, indicate that average concentrations of tributyl
tin (TBT) in polychaetes collected from within the HARS are approximately twice that of
polychaetes collected from the reference site. These data suggest enrichment of TBTs in sediments
from the HARS. Moreover, data on concentrations of TBTs in surficial sediments taken from
throughout the New York/New Jersey Harbor indicate that dredge materials from these areas may
be a significant source of TBTs to the HARS (EPA 1998).
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While the concentrations of TBTs in surficial sediments of the New York/New Jersey Harbor are
variable, (i.e., < detection limits to > 500 ppb), they appear to show consistent spatial patterns.
These data suggest that it may be possible to identify sediments that would not require testing for
TBTs. Moreover, the concentrations of TBTs in these polychaetes are generally low and, given the
differences in organic carbon content and grain size between the HARS and reference sites it is
difficult to determine the significance of this modest apparent enrichment of TBTs.
Given the spatial variability and relatively low concentrations of TBTs in benthos, I would propose
that the party seeking the permit for disposal of dredged materials in the HARS be given the option
of developing lines of evidence as to why TBTs should not be included in the bioaccumulation
testing. Information to be considered could include: 1) existing data on TBTs in sediments to be
dredged; 2) historic and ongoing anthropogenic activities in the vicinity of the sediments; 3)
sediment characteristics (e.g., organic carbon content etc.); and 4) arguments based on TBT
partitioning in the environment (Meador, 2000).
Proposed Addition to the Analvte List: Coplanar PCB congeners
9. Inclusion of dioxin-like PCBs if the approach for evaluating dioxins is modified:
If EPA Region 2 elects to evaluate dioxin-like toxicity based on the 2,3,7,8-TCDD toxic
equivalency (TEQ) approach, it should include data on coplanar PCBs in that evaluation. These
data would provide a basis for a more complete evaluation of potential for toxicity attributable to
those compounds that act via the AH receptor pathway.
I am unclear as to how to address the second part of this question, (i.e., "If so, how?") I see a
number of potential questions here ranging from: 1) how the analysis of coplanar PCBs should be
conducted; to 2) how the coplanar PCBs should be incorporated in a human health risk assessment.
As I am not a chemist and do not normally conduct human health risk assessments, I do not feel
qualified to address these issues.
Comparison to Reference
10. Assigning values to tissue residues that are less than detection limits in statistical
comparisons with the reference site:
The current policy for dealing with data that are below detection limits treats data from the HARS
and reference site equally when the detection limit is below the method detection limit (MDL)
(e.g., data from both sites, that are below detection limits, are given the value of lA the MDL).
However, when the detection limit is above the MDL the data for the reference site and HARS are
treated differently. In this situation, the non-detect data from the reference are assigned a value of
0 while that from the HARS is reported as the detection limit. This approach, in effect, penalizes
3/11
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the permitee for high detection limits by increased the probability of finding statistical differences
between the HARS and the reference site.
Although I have no problems with the general approach, the discussion in this document is
unnecessarily indirect and obscure. As an example, no rationale is presented for this approach nor
are the consequences of the policy made explicit. If the concern is the increased uncertainty
associated with a higher detection limit, it should be so stated. Moreover, it should be made clear
to the permitee that the goal is to obtain a detection limit that is less than the MDL for each analyte
considered. It should also be made clear that although analyses with detection limits greater than
the MDL will be accepted, the data will be interpreted in a way that increases the probability of
finding statistical differences between the HARS site and the reference site.
11. Use of functional groupings in statistical comparisons to reference site:
In theory, where appropriate data are available, it is appropriate to use functional groupings to
evaluate bioaccumulation and toxicity for the grouping. However, uncertainties due to variations
in composition of individual compounds within the groups must be acknowledged and accounted
for in the methods.
In addition, it is not clear how data that are less than detection limits are to be addressed for the
individual analytes that will be summed for the functional groupings. The HARS Framework is
silent on this issue. The March 14,1997 EPA/CEAN document indicates that for total PCB and
DDT "conservative estimates" of concentrations are to be used for constituents that are below
detection limits but no details on specific methods are presented. Do the rules presented in the
preceding paragraph of the HARS Framework apply to individual PCB congeners or PAHs? If so,
then the issues I raised in my previous comment apply here as well. If not then what is the
rationale for treating these analytes differently?
12. Use of multiplier based on log K«w in adjusting for steady state:
My primary concern with the use of the log KoW is that this approach doesn't take into account the
potential for metabolism of organic compounds such as the PAHs. If significant metabolism is
occurring the use of a log KoW based steady state correction factor could substantially overestimate
the steady state concentration (Baussant, et al., 2001). However, it is unclear how significant this
issue is given that invertebrate species often have limited capacities for metabolizing organic
compounds. The uncertainty introduced by this approach should be addressed based on the
available literature regarding metabolism of organic compounds by polychaetes and bivalves.
13. Use of parent correction factor for alkylated PAHs:
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As alkylated PAHs are generally more hydrophobic than the parent compounds, they will have a
correspondingly higher log Kow- The application of the correction factor associated with the parent
PAH has the potential of systematically underestimating the Kow for the alkylated derivative and
thus the predicted steady state. However, without specific KoW data for the alkylated PAHs it is
unclear if the differences in the K<,w relative to the parent PAHs would be sufficient to result in a
modification of the proposed multiplier. This issue could be evaluated to determine if it represents
a significant source of error in the analysis. However, while the inclusion of alkylated PAHs
significantly increased exposure estimates, there are little or no data on the relative carcinogenicity
and non-cancer toxicity of these additional compounds. Therefore, clarifying the uncertainties
associated with the use of the correction factor for the parent compound may not warrant a
significant effort.
14. Use of M. nausta data for estimating steady-state multipliers for N. virens for DDTs and
dieldrin:
Steady state data for total PCBs are available for both M. nausta and N. virens. These data indicate
that for PCBs, M. nausta archives steady state significantly faster then does N. virens. This
difference is reflected in the lower multiplier for M. nausta (Table 4). If we assume a similar
relationship exists for DDT derivatives, than the use of a M. nausta multiplier would tend to
underestimate the steady state concentration of DDT derivatives in N. virens.
15. Adequacy of approach for establishing steady state:
Other than the issues addressed in comments on Charges 12,13 and 14, the approach taken to
adjust contaminant bioaccumulation data are adequate and generally consistent with the literature.
16. Sources of uncertainty in estimating steady state:
I have addressed what I believe to be the specific sources of uncertainty in my responses to
Charges 12 through 14.
Adjustment to Steady State: Metals
17. Methodology used in deriving a steady state multiplier for non-essential metals:
The multiplier for non-essential metals is not based on steady state considerations. Instead, Data
on the range of variability in concentrations of non-essential metal (e.g., silver, cadmium, mercury
and lead) in tissues of organisms from the bight are used to develop a safety factor. Because there
was no more than a three fold variation in concentrations for any of the non-essential metals, a
safety factor of 3 was adopted for all of the metals.
5/11
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The approach is conservative in that it assumes implicitly that the tissue data from the bioassays
reflects the lowest concentrations likely to be encountered in the field then multiplies those data by
a safety factor of 3 to estimate a potential worst-case scenario. Moreover, the actual min/max
ranges for silver, cadmium, mercury and lead are all less than 3 (e.g., 2.3 to 2.8) adding to the
conservatism. Due to the multiple layers of conservatism, consideration should be given to using
the actual min/max ratios for each metal instead of applying a generic factor of 3.0 to all of the
metals (See data in Appendix G). Alternatively a probabilistic approach might be used here in
place of a deterministic point estimate.
Human Health Evaluations: Overall
18. Appropriateness of factors used in equations for the human health evaluation:
My general concern is that parameters used in estimating potential for exposure of humans to
HARS related chemicals are based on conservative point estimates and do not adequately account
for the uncertainty and variability (see comments on Charge 1).
I have also provided specific comments on several of these factors in my response to other
Charges. These include:
• The use of parent compound potency factors for alkylated PAHs - Charge 23.
• The whole -body to fillet conversion factor - Charge 20.
In evaluating the derivation of the site use factor, I am concerned that it only takes into account
temporal variability and not spatial variability. This could be addressed by evaluating the foraging
ranges, relative to the HARS, of species that contribute most to the existing site use factor (e.g.,
flounders, cod, blue fish and striped bass).
The methods used to develop trophic transfer factors appear reasonable and the factors presented in
Table 5 are generally consistent with data from published field studies. However, I could find no
technical support for the use of a trophic transfer factor of 1 for chromium, lead and silver. A
trophic transfer factor of 1 will likely result in a substantial overestimate of potential for these
metals to accumulate in fish tissues. Consideration should be given to reevaluate the trophic
transfer factor for these metals.
As I have had limited experience with human health evaluations I do not feel qualified to address
the remaining assumptions (e.g., potency factors, seafood consumption rates and exposure
duration).
19. Methods used to address human health exposure parameters.
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I have commented on most of the exposure factors in my response to other Charges. These
include:
• The whole -body to fillet conversion factor - Charge 20.
• The site use factor - Charge 18
• The trophic transfer factors - Charge 18
As I have had limited experience with human health evaluations I do not feel qualified to address
seafood consumption rates or exposure duration.
20. Conversions of fish whole body data to fillet data:
Lipophilic compounds:
Appendix K proposes that a single whole-body to fillet ratio of 1.35 be used for all lipophilic
compounds and all fish species. At steady state, lipophilic compounds will partition among tissues
of the fish based on the lipid content of those tissues. The proposed approach assumes that the lipid
content of the fillet and whole body are consistent among fish species. However, the relative lipid
content of tissues can vary substantially among fish species. Appendix K indicates that proposed
ratio was derived from fish from the New York State and the Great Lakes. It is unclear which
species were used in deriving the proposed ratio or how lipid content of the tissues of these species
might compare with those that forage in the area of the HARS. The uncertainties introduced by the
use of a single ratio, and their potential impact on the evaluation of risk, should be evaluated and
explicitly addressed in this document.
Inorganic compounds:
Whole-body to fillet ratios for arsenic, chromium and mercury are derived from regression analysis
of data from black bass (Bevelhimer et al., 1997). The underlying assumption in using these ratios
is that the species and exposure range from this study are representative of the species of concern
and exposure ranges found at the HARS. These assumptions should be explicitly addressed and
any uncertainties acknowledged.
Appendix K proposes to use a whole-body to fillet ratio of 1 for the remaining metals (Ag, Cd, Cu,
Ni, Pb and Zn). No justification is presented for a ratio of 1 for these metals. Moreover, for a
metal such as Cd a ratio of 1 does not accurately reflect the relationship between the concentrations
in the whole-body and fillet. As an example, data presented in Handy (1992), indicate a whole-
body to fillet ratio for Cd of about 2 in control fish. In addition, subsequent exposure to cadmium
resulted in an eight-fold increase in whole-body concentrations but no significant change in fillet
concentrations, giving a whole-body to fillet ratio of 16. Similar but less dramatic changes were
seen for the whole-body to fillet ratio for copper following episodic exposure (Handy, 1991).
These data, and similar data from numerous other studies, suggest that fish are able to regulate
tissue-level compartmentalization of metals over a range of exposures and indicate that a single
whole-body to fillet ratio of 1 does not accurately reflect this relationship within a fish over time or
7/11
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among species. In most instances using a ratio of 1 to extrapolate whole-body to fillet will likely
overestimate concentrations in fillets. However, given the ability of fish to regulate metal
compartmentalization the degree of overestimation is difficult to predict. As with lipophilic
compounds, the uncertainties introduced by the use of a single ratio, and their potential impact on
the evaluation of risk, should be evaluated and explicitly addressed.
21. Focusing on key fish species in the HARS:
Focusing on key species would provide a more environmentally realistic basis for evaluating
exposure than the current approach. Factors to be considered in selecting key species might
include: 1) relative importance for both commercial and sports fishing in the vicinity of the
HARS; 2) a foraging range that could result in significant time spent foraging within the HARS;
and 3) availability of regional data on other factors that will be important the analysis (feeding
habits, size distributions, lipid content in whole body and fillet etc.).
22. Use of the most toxic metal form in evaluating bioaccumulation, trophic transfer and
risk:
For chromium and mercury, the use of potency and exposure factors for the most toxic valence
species or form of the metal will most likely overestimate risk in a marine environment such as the
HARS. This conservative approach is justified given the uncertainty resulting in the absence of
data on speciation of chromium and mercury. Moreover, the policy allows proponents of dredging
projects to provide data on metal speciation to reduce this uncertainty.
For arsenic, the use of a factor of 0.1 will likely still overstate the actual concentrations of
inorganic arsenic in bioassay species. Again this conservatism is appropriate in the absence of
data on arsenic speciation. However, as with chromium and mercury, proponents of dredging
should be given the opportunity of providing data on arsenic speciation rather than relying on a
generic application factor.
23. Estimating the potency of alkylated PAHs based on that of the parent compound:
I am not aware of data that could be used to evaluate the potential cancer or non-cancer potency of
alkylated PAHs in an evaluation of human risk. Therefore, I have no basis for comparing the
relative potencies of alkylated PAHs and that of their parent compounds. I believe the real
questions here are: 1) Do we ignore the significant mass of PAHs represented by the alkylated
compounds in our exposure analysis? or 2) Do we include this additional mass in our estimates of
exposure and guess as to its potency? As the alkylated PAHs often represent the substantial
majority of PAHs in the sediments (Irwin et al., 1997), not accounting for these compounds has the
potential to substantially underestimate the total risk attributable to PAHs. Therefore, I favor
including the alkylated PAHs in the analysis and, in the absence of data, extrapolating from
potency data from the parent compound. The uncertainty associated with this extrapolation should
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be addressed in the Framework Document and taken into account in evaluating potential risk to
humans.
24. Potential for human exposure to PAHs through consumption of finfish and other
seafood:
My general concern is that each of the parameters used in estimating potential for exposure are
based on conservative point estimates. These point estimates do not adequately account for the
uncertainty and variability associated with each parameter. Consideration should be given to
employing a probabilistic approach.
In addition the evaluation of human exposure to PAHs through the food chain should take into
account the ability of the various taxonomic groups to metabolize these compounds. On this basis,
the potential for exposure to PAHs via ingestion of finfish should be very limited given their ability
to metabolize these compounds. However, exposure to PAHs via ingestion of invertebrate species
would likely vary substantially depending upon their capacity to metabolize these compounds.
These differences among taxonomic groups can be addressed based on available data in the
literature.
25. Sources of uncertainty in approaches from Section E.
I have commented the uncertainties with a number of the approaches presented in Section E in my
response to other charges. These include:
• The whole -body to fillet conversion factor - Charge 20.
• The site use factor - Charge 18
• The trophic transfer factors - Charge 18
As I have had limited experience with human health evaluations I do not feel qualified to address
uncertainties associated with assumptions regarding seafood consumption rates or exposure
duration.
26. Evaluating the toxicity of organotins:
I have had limited experience with human health evaluations and I do not feel qualified to address
this issue.
27. Use of alkylated and parent PAH data in evaluating risk:
I have commented on the appropriateness and uncertainties associated with the proposed approach
for converting and using the analytical data for alkylated and parent PAHs in my responses to
previous charges (see responses to Charges 2,4, 5,12,13,15,23,24).
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Human Health Evaluations: Comparisons to HARS-Specific Values
28. Use of disaggregate modeling:
I have had limited experience with human health evaluations and I do not feel qualified to address
this issue.
Human Health Evaluations: Consideration of Combined Effects
29. Estimating combined effects mixtures on carcinogenicity and non-cancer impacts:
I have had limited experience with human health evaluations and I do not feel qualified to address
these issues.
30. Conceptual model for evaluating fish exposures to dredged materials in the HARS and
human exposure through the ingestion of seafood:
The proposed framework does not explicitly present a discussion of a conceptual model. However,
the major components and assumptions of the framework constitute an implicit conceptual model.
In reviewing these assumptions I have a general concern that I have stated previously. The major
parameters in the model are represented as conservative point estimates and no attempt is made to
quantify the uncertainty and variability. EPA should consider a probabilistic analysis to better
quantify uncertainty and variability that are implicit in the proposed framework (see comments on
Charge 1).
I am also concerned that the reference site is inappropriate given the higher grain size and lower
organic carbon content of reference sediments, relative to HARS sediments and most dredged
materials. The use of this reference site will increase the potential for statistical exceedances and
will limit the usefulness of this initial step in the evaluation.
Specific issues regarding other aspects of conceptual model have been addressed in responses to
previous Charges. These assumptions are discussed briefly below:
References:
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Baumard, P., Budzinski, H., Garrigues, P., Sorbe, J.C., Burgeot, T., and Bellocq, J. 1998.
Concentrations of PAHs (polycyclic aromatic hydrocarbons) in various marine organisms in
relation to those in sediments and to trophic level. Marine Pollution Bulletin 36(12):951-960.
Bevelhimer, M.S., J.J. Beauchamp, B.E. Sample, and G.R. Southworth. 1997. Estimation of
whole-fish contaminant concentrations from fish fillet data. U.S. Department of Energy,
ES/ER/TM-202.
Handy, R.D. 1992. The assessment of Episodic Metal Pollution. 11. The effects of cadmium and
copper enriched diets on tissue contaminant analysis in rainbow trout. Arch. Environ. Contam.
Toxicol. 22: 82-87.
Irwin, R.J., VanMouwerik, M., Stevens, L., Seese, M.D., and Basham, W. 1997. Environmental
Contaminants Encyclopedia. National Park Service, Water Resources Division, Fort Collins,
Colorado.
Meador, J.P. 2000. Predicting the fate and effects of tributyl tin in marine systems. Rev. Environ.
Contam.Tocicol. 166:1-48.
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Comments Received from
Peter F. Landrum
March 15,2002
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Review: Proposed Bioaccumulation Testing Evaluation Framework for Assessing
the Suitability of dredged Material to be placed at the Historic Remediation Site
(HARS).
By
Peter F. Landrum
Great Lakes Environmental Research Laboratory, NOAA
2205 Commonwealth Blvd.
Ann Arbor, Ml 48105
Charge 1: Throughout the proposed process, there are various
uncertainties introduced. Please identify key areas of uncertainty that need
to be addressed. Are there additional data sources or parameters that
could be used to address these areas? What methods are available for
describing and accounting for these uncertainties in the calculation of
HASR-Specific Values? Of the methods available, which would you
recommend for consideration and why? Please consider the implications
of implementing these methods in the regulatory framework. Please
include and evaluation of probabilistic and deterministic methods in your
discussion.
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One of the great uncertainties in the evaluation of sediment is the absence
of use of longer-term bioassays. The limitation of using only 10 d toxicity tests as
the fundamental toxicity test to evaluate the potential toxicity of disposal material
leaves a gap in the protection of the disposal site from sediments that may result
in long-term effects. In freshwater systems, there is a move to incorporate
longer-term bioassays, e.g., 28 d Hyalella growth and mortality and 40-d
reproductive tests. These test designs have been incorporated into the latest
version of the EPA Guidance for measuring the toxicity and bioaccumulation of
sediment-associated contaminants with freshwater invertebrates (USEPA 2000).
It certainly seems that considerations of longer-term bioassays such as that by
Gray et al. (1998), which is a long-term toxicity test with Leptocherius, should be
part of the methods for protecting the aquatic life at HARS. Limiting the toxicity
testing to just a 10-d test is equivalent to only evaluating toxicants with an acute
toxicity test. Chronic tests are critical to understanding the full potential of the
toxicity of sediment-associated contaminants.
The material chosen as reference limits the evaluation of contaminant
bioaccumulation from proposed dredged material. Selection of the reference site
is critical for proper evaluation of the bioaccumulation bioassay. No single site
can be presumed to be a perfect reference. This question was considered
crucial in a previous evaluation the bioaccumulation bioassay for dredged
material requested by EPA region 1 in 1994 (Metalf and Eddy 1995). One of the
important conclusions from that workshop was the establishment of a regional
tissue background value for each of the contaminants of interest against which to
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compare the bioaccumulation test. This was thought to be much more robust
that selecting a single sediment or single site for the evaluation. EPA Region 1
collected the data (tissue levels) from all reference sites in the region. This
allowed construction of the range of background concentrations in the region and
they were ranked. Then bioaccumulation test data could be compared to the
concentrations say below which 95% of all background values fell. This provides
a more robust reference than a single site. The decision on where to make the
cutoff for reference values can be selected based on the needs of the program.
The data to produce a similar evaluation for Region 2 already exists from the
monitoring and operational studies that have already been performed. It would
only be necessary to gather the data and rank them for the reference and
background sites in the region.
Several specific models are used in this process to develop deterministic
results without any indication that model uncertainties have been addressed.
One of the ways to reduce the uncertainties throughout this process is to look
closely at the models and determine the expected range of response of the
models based on the variability in the parameters used to set the deterministic
solutions. Specifically, there is a tendency in the document to use the data from
only one or two studies to parameterize the models. There are areas where
regionally specific data might be derived and provide improved indications of the
variation. This data may well be available from past operational data. For
instance, previous 28-d bioaccumulation studies have been performed with
materials that came out of the region and were dumped at the original site.
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Unfortunately, monitoring data for benthic organism tissue with synoptic sediment
concentrations at the dumpsite are apparently not available (according to EPA
per phone conversation with N. Bonnevie). If these data were available, a
regionally based factor relating the performance of the 28-d study to in situ
steady state could be established rather than relying on factors from one or two
specific studies that only examined a single sediment. However, there are
sediment concentrations and polychaete tissue concentrations that have been
collected from the same site although not synoptically. These could be used to
make a first cut at a regional specific relationship between the 28d bioassay and
in situ conditions. These data could confirm the current multipliers or indicate a
need for improvement. The polychaetes were not depurated and this is not really
an issue for the organic contaminants, as past work has shown that depuration
has little impact on the concentration of organic contaminants in oligochaetes
(Oliver 1985). The lack of depuration may be significant for the metals, however.
The time difference between the collection of the sediment measurements and
the polychaete measurements would likely have minimal impact since the
sediment concentrations would not be expected to change rapidly. The
approach of comparing in situ data to the 28-d bioaccumulation response would
be an improved manner for establishing HARS specific values. There may well
be other monitoring data in the region (not specific to the HARS) that would be
appropriate for making the comparison between the 28-d bioaccumulation assay
and the in situ exposures. The approach for metals has been half-way done
based on the Battelle 1996,1997 studies that compared metal concentrations in
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sediment and in tissues for samples from the site (See page 19 of framework).
All that remains is to make comparisons to 28-d bioaccumulation studies for
comparison (See details below).
It would also be useful to develop a review of tissue concentrations for
benthic organisms for the region against which to make comparisons. If the
addition of dredged material is truly intended to remediate the site, one measure
of progress would be the extent to which the contamination was predicted to
approach background concentrations. EPA Region 1 (Metcalf and Eddy 1995)
did a review of the background concentrations within the region from studies that
had measured concentrations from reference or uncontaminated sites. The data
may well be relevant for this region as some of the data came from the New York
area. These background tissue concentrations were ranked so that they could
be compared with the potential accumulation in bioassays, potentially
contaminated sites, or for sites where remediation had occurred. These data
provide a clear delineation of which areas are essentially at the regional
background level and which are clearly above. Since the data are on tissue
levels for benthic organisms, the 28-d bioassay corrected either with the current
multipliers or a more appropriate regional multiplier could be evaluated to
determine whether the sediment would lead to levels that are currently clearly
above background. This approach should be particularly useful for the HARS
site where there is likely some expectation that the sediments will be remediated
and the potential of a particular sediment to help attain that goal could be
evaluated. Comparing to regional values removes uncertainty and helps insure
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that one is not being too strict or lenient in the concentration allowed into the
HARS site.
As often as the charges to the peer review team have asked the team to
point out uncertainties, it is curious that no formal uncertainty and sensitivity
evaluation of the various models was attempted in formulating this document.
This seems to be one of the major oversights of the document.
Deterministic versus probabilistic models: Both approaches have
advantages and disadvantages. The probabilistic models more readily allow the
presentation of the uncertainty of the models, resulting from both the uncertainty
and variability of the values parameterizing the models, to be expressed in the
output. However, they require a large amount of data. Further, they are harder
to troubleshoot for problems in model structure. They certainly allow for
decisions that do not have to be made as black or white. Deterministic models
make the evaluation of model structure easier so that model error is more readily
detectable. The uncertainties of the model can be developed through the use of
Monte Carlo simulation to show the impact of parameter variability on the results.
In the case of the HARS site and the evaluation of bioaccumulation, it seems that
the situation is severely data limited, thus there is not a good justification to
proceed to a probabilistic modeling framework. However, should the data
become available for developing a probabilistic model in the future, the
appropriate model could be incorporated into the regulatory environment.
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The models used in this process for decisions should undergo a rigorous
uncertainty and sensitivity analysis. This includes sub-models that are used to
set parameter estimates e.g., McFarland 1995. Further, the assumptions of the
models should be clearly laid out and the potential impact of these assumptions
evaluated. Most models have explicit assumptions that the modeler attempts to
address. In addition most models have implicit assumptions that are often
overlooked. An effort should be made to insure that both types of assumptions
are addressed. Further, the structure of the model can lead to inaccuracy of
prediction. The model used most in this work is that of Gobas 1993. An
examination of the structure of the model and its ability to reflect the food web of
interest is critical to insure that there was correct parameterization for the current
use. The model was presumably structured specifically for the characteristics of
the regional food web, but this was not communicated within the document.
There was described in the document the use of a simplified food web for use in
the modeling. This may well be a limitation to the accuracy of the predictions.
For instance, in freshwater lakes, the length of the food web was crucial to
predicting the bioaccumulation of PCBs in upper trophic levels (Rasmussen et al.
1990). Thus, the food web model should reflect the existing food web as closely
as possible. This might well be established using a stable isotope approach to
insure that the length and structure of the food web is accurate as possible. This
approach can also be used to develop empirical bioaccumulation factors that
may well be more accurate than computer models (e.g., Broman et al. 1992).
Based on the reference EPA 1995, which is used for developing multipliers, there
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is no indication that an uncertainty analysis was performed or that the food web
structure was examined critically. It is important that both uncertainty and
sensitivity analyses of all of the models used in the process are performed and
that the model structures be examined to insure that they accurately reflect the
food web structure of interest. It is important to demonstrate that the
deterministic solutions employed for this process will lead to protective solutions
as represented by the confidence intervals from the uncertainty analyses.
Charge 2: Is measurement of the 16 priority pollutant PAHs (i.e., parent
PAHs) sufficient for characterizing the risks associated with the total PAH
bioaccumulated by organisms exposed to dredged material proposed for
placement at the HARS? Does measurement of the alkylated compounds
significantly improve risk assessment of PAHs? Literature exists suggesting
that incorporation of the alkylated PAH is important for full evaluation of PAH
toxicity for the non-polar narcosis (anesthetic) mode of action (DiToro and
McGrath 2000). This is based on the notion that the compounds will act
additively to produce non-polar narcosis. It is clear that alkylated PAH can
produce toxic effects including carcinogenicity. For instance,
dimethylbenzanthracene is a well-recognized potent carcinogen. Further, some
of the alkylated PAH demonstrate the potential to be photoactivated and produce
photo-induced toxicity (Bose et al. 1998). Thus, if bioaccumulated they will be of
concern and are at the least expected to contribute to the non-polar narcosis
based on the work of DiToro and McGrath (2000). It is further expected that if
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some of the recognized carcinogens such as dimethybenzanthracene are
accumulated up the food web cancer risk would occur. From a bioaccumulation
perspective the compounds should be added. Failure to include the alkylated
PAH could result in lower estimated risk than would actually occur from
contaminated sediments.
Charge 3: Is the proposed adaptation of EPA Method 8270 (Appendix D)
acceptable and appropriate for regulatory decision-making? If not what is
an acceptable and appropriate method? The method in appendix D had
sufficient detection limits and precision for the analysis of the PAH including
alkylated PAH. However, the analytical standardization of the method is all
based on the parent compounds (per Table 2). It would seem that such a
method should include some alkylated PAH for surrogates, internal standards,
and matrix spikes to insure that the method was performing well for the alkylated
PAH.
Charge 4: Under what specific conditions would the testing for alkylated
PAHs for a particular project be appropriate and warranted? Alkylated PAH
tend to dominate the PAH profile of petroleum sources, e.g., Wang et al. (1999),
while parent PAH tend to dominate anthropogenic combustion sources (from
Gabos et al. 2001). Since alkylated PAH can greatly dominate the PAH profile in
those cases where petroleum is a major contributor of the PAH, under these
conditions, the particular focus should be placed on the alkylated PAH. This is
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not to indicate that the alkylated PAH should be neglected for the remainder of
the time. They likely should be included for all studies and included as
contributors to the PAH toxicology.
Charge 5: What uncertainties would be introduced within the analysis of
risk should alkylated PAHs be included? What steps could be taken to
account for these uncertainties in decision-making? Given the likelihood
the method for using non-detects as described in EPA/CENAN (1997) will
result in an overestimate of risk, what are the implications? Alkylated PAH
have not been as thoroughly studied as have the parent compounds. This
includes their mechanisms of action. Thus, including the alkylated PAH into a
risk analysis could be in error depending on the mechanism of action being
considered. The absence of sufficient study includes bioaccumulation, food web
transfer, and bioavailability issues as well as fundamental toxicology. This is an
area where additional research could greatly help improve our understanding.
Incorporating the alkylated compounds in a risk assessment would be
conservative. Leaving the alkylated PAH out of a risk assessment would likely
leave substantial gaps and result in an underestimated risk.
Charges 6,7,8: These are outside my area of expertise.
Charge 9: If the approach for evaluating dioxin is modified should it include
the contribution of coplanar PCBs with dioxin-like activity as proposed? If
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so How? The coplanar PCBs should immediately be added to the contribution
of dioxin-like activity. The literature shows that coplanar PCBs may well produce
more dioxin-like activity in environmental samples than the dioxins and
dibenzofurans because of their relatively higher concentration and higher
bioavailability despite their lower compound specific activity at the receptor (e.g.,
Ludwig et al. 1996). Using a TEF approach, it is now possible to incorporate the
role of coplanar PCBs in dioxin-like activity for both fish and wildlife (Tillitt 1999).
Charge 10: Please consider the policy for assigning values to tissue
residues that are reported as "
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multipliers for steady state was only derived for PAH and not for other
compounds classes. Therefore, it should not be applied beyond the class of
contaminants for which it was developed. There is ample data in the literature
demonstrating that chlorinated hydrocarbons are more bioavailable than PAH of
similar log KoW (e.g., Landrum et al. 1997) and so the values that would be
calculated from this figure could easily be inappropriate. Further, the method to
derive the figure has not been peer reviewed and it was not clear which organism
was used for the determination. It would be superior to establish regional
multipliers that could be applied to the HARS. As suggested above, data exists
for all of the 28-d bioassays that have been performed for the dumping of
previous material at the site. Further, if synoptic sediment and tissue
concentrations were determined for benthos at the site then regional multipliers,
perhaps adjusted for sediment composition such as the amount of organic
matter, could be developed. The monitoring data for the region should be
reviewed to establish the regional multiplier relationship adjusting for sediment
composition and perhaps organism species.
Charge 13: Given the increased hydrophobicity of alkylated PAHs, is the
use of the correction factor associated with corresponding parent an
appropriate approach for estimating steady state residues of alkylated
PAHs? If not, please elaborate. If the alkylated PAH are more hydrophobic than
the parent PAH, then it is more appropriate to use a correction factor that is more
in line with the respective log Kow and not necessarily that of the parent
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compound from which the alkylated compound is derived. The reason for such a
choice is that the hydrophobicity of the compound is generally recognized as the
driving force for the bioaccumulation and partitioning behavior of the molecule.
Therefore, to model alkylated PAH based on the lower log Kow of the parent
compound would not seem to be appropriate. However, a QSAR approach
would work, assuming there was a relationship between log Kow and the
correction factor for the parent PAH then that relationship with the appropriate log
Kow for the alkylated PAH could be used.
Charge 14: For the DDT derivatives and dieldrin, please comment on the
appropriateness of using M. nasuta data rather than N. virens specific data
in the estimation of steady state multipliers. It appears that this document
makes too much use of a single study by Lee et al. (1994) to set the multipliers.
This seems to be one of the weak links in the process. Lee et al. (1994) used
sediments from only one site for the long-term toxicokinetics study. Thus, the
results represent the bioavailability characteristics of a single sediment.
However, it has been demonstrated over and over in the literature that the
bioavailability varies substantially with changes in sediment characteristics (e.g.
Landrum et al. 1997). Thus, choosing one study to base the multiplier seems
totally inappropriate. It would be better to mine existing data or develop data to
establish the multiplier between the 28 d study and steady state. Studies taken
from regional sediments and compared to the bioaccumulation observed at the
dumpsite would be a far superior approach to establish the relationship between
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the 28 d bioaccumulation and the actual in situ steady state. The development of
a regional correction factor based on the sediments from the region would also
allow for the variability in the multiplication factor to be established. This is one
area of uncertainty that could be established for the process. Based on the
above, I would not use either the M. nasuta or the N. virens data. I am
particularly suspicious about the value for p.p'-DDT established with M. nasuta.
This value seems out of line with the other compounds of similar log Kow.
Unless there are other data to support the large multiplier value for DDT then it
should be re-established. I suspect that the low accumulation of DDT relative to
the sediment could well result from biotransformation of the DDT to DDE during
the test. This could not be ruled out in the study that was performed to establish
the multiplier.
Charge 15: Are the approaches taken to adjust organic contaminant
bioaccumulation data to steady state adequate? Do the proposed
multipliers agree with previously published studies (i.e., to they appear
reasonable)? If not, please elaborate. See comments above for generating a
regional multiplier. Also, it would seem that even if the data are not available in
the region, there are likely many sites throughout the US where monitoring data
could be compared to data generated with the 28-d bioassay data. This
empirical approach would be superior to the use of individual studies.
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Charge 16: What are the major sources of uncertainty associated with the
approaches? What alternative approaches would reduce the
uncertainties? How could these uncertainties be described and accounted
for in decision-making? One major uncertainty in the approach is establishing
the multiplier between 28 d and steady state. This could be better established
through development of a regional multiplier using the data that is already
available in the region. This would certainly be more appropriate than using the
data from a single study to establish the multipliers. The relationship between
the 28 d results will be compound, sediment, and organism dependent. Thus, it
makes good sense to establish a regional value. The advantage of establishing
the relationship between 28 d studies and the observed steady state from in situ
organisms is that similar multipliers could be established for metals. This would
help reduce the uncertainty of the approach.
For that matter, it is possible to establish regional links between the
sediments and fish or lobsters (an exposure route that is present but not
addressed in the bioaccumulation process). These empirical relationships do not
require a model and could provide the range of bioaccumulation values that are
reasonable for the region. The set of assumptions that would need to be made
could be agreed on and the existing data would likely be available to establish
the relationships. Any approach that could use actual data to establish the
relationship between the sediments and the bioaccumulation would reduce the
uncertainty of the risk evaluation. Any time a model is employed to evaluate the
expected food web bioaccumulation relationship the uncertainties will be large.
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Finally, if the variability and the uncertainties in the parameters are not
known and not evaluated in the models, the extent of uncertainty of the models
remains unknown. Thus, it is not clear that the values chosen will lead to the
levels of protection desired. The uncertainties of the models resulting from the
variability in the parameters must be evaluated. This proposed uncertainty
analysis must be done as a minimum for the proposed framework to be used.
Charge 17: In your opinion, is the methodology (see Appendix G) followed
to derive the steady state multiplier for non-essential metals (i.e. a factor of
three) scientifically appropriate? Please elaborate. Do you have any
recommendations of additional or alternate methodologies or information
that can be used to either supplement or replace the proposed method?
This methodology did not attempt to account for the characteristics of the
sediment and instead applies a safety factor without establishing the relationship
between the 28 d bioaccumulation study and the values found in situ. To use
this approach, the relationship between the tissue concentrations in 28 d
bioaccumulation studies should first be compared to the values from the region.
If the relationship is essentially one, that is the 28 d data are not statistically
different from the regional values, then the factor of 3 is conservative and
appropriate. If however, the ratio between mean of the 28 d tissue
concentrations and the regional tissue concentrations is less than one then an
additional factor will be required. The EPA did half of the work now all that is
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required is to compare to the data from previous bioaccumulation tests to the in
situ results.
Charge 18: Please comment on each factor listed above (and in Table 5) as
to its appropriateness for use in the equations listed above. Would you
recommend additional factors? Would you change of modify the equations
as written above? The only factor that I can comment on is the trophic transfer
factor. The trophic transfer factor between algal and zebra mussels is much
larger than 0.1. This transfer was 0.53 for BaP (Bruner et al. 1994) and 0.95 for
chlorinated compounds (Ma et al. 1999). Further, the transfer between BaP and
sediment for benthos has been found to range from 0.2 for zebra mussels
(Bruner) to 0.5 for Diporeia (Lydy and Landrum 1993). Thus, selecting 0.1 for the
first link is probably low. The remainder of the transfers may well be limited for
parent PAH because the extent of biotransformation increases with increasing
trophic transfer. Further, all benthic organisms are not exposed equally (Harkey
et al. 1994, Wilcock 1993), suggesting that the specifics of the food web model
may lead to either elevated or lower exposure in the trophic food web depending
on the specific trophic links. There have been studies that have established the
food web links through the use of the ratios of stable isotopes of nitrogen (e.g.
Broman et al. 1992). It would be useful to establish the specific food web links
for the various fish of opportunity to make certain that the models used to derive
the trophic transfer are accurately constructed.
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It seems that the trophic transfer coefficient for chlorinated compounds is
low compared to that observed in freshwater where the magnification is about a
factor of 3.5 with each trophic link (e.g., Rasmussen et al. 1990). These data
were generated from data in lakes and not by a model. It would seem that
similar data may exist for the region and better trophic transfer factors could be
obtained. The use of model results to create trophic transfer factors creates
uncertainty. Making empirical measurements with the food web position
established through the use of stable isotope data would provide improved
information on variability and permit quantitation of the impact of the uncertainty
in the calculation. Currently the uncertainty is hidden in the deterministic solution
from the model. At the vary least, the model uncertainty in the trophic transfer
should be determined and explicitly provided.
DDT and its degradation products should not be treated as a group. First,
the compounds, DDT, DDD, and DDE are not equipotent (e.g., Lotufo et al.
2000). Second, they all do not produce toxicity by the same mechanism of
action. DDT is recognized as a neurotoxin (Gilman et al. 1980) while DDE is
more widely recognized as affecting reproduction (Kelce et al. 1995) and based
on the data from Lotufo et al. (2000) DDE appears to act as a non-polar narcotic
for the purposes of evaluating mortality. Third, they do not all have the same
physical/chemical characteristics. Thus, to treat the mixture as a group is
inappropriate toxicologically.
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Again, I suggest that a formal uncertainty and sensitivity analysis be
performed on the risk models to insure that the deterministic solution is providing
the level of protection expected.
Charges 19,20,21,22: These charges are outside my area of expertise.
Charge 23: Is the assumption that the potency of alkylated PAHs can be
estimated by the potency of the parent PAH appropriate? Is this
assumption likely to result in an under- or over estimate of the risk
associated the alkylated PAH? No, it is not appropriate to assume that the
potency of the alkylated PAHs can be estimated from the potency of the parent
PAH. This would be an appropriate assumption for non-polar narcosis but would
be out of line for estimating the potential for carcinognicity. For instance,
methylbenzanthracene derivatives are more carcinogenic than the parent
(Newman 1986) and further the position of the methyl group alters the
carcinogenicity (from Yang et al. 1981). Similar differences might well be
expected for any mode of action except non-polar narcosis.
However, these compounds have not been well studied and so the overall
use of the potency of the parent to reflect the alkylated compound may be the
only approach available. If there was sufficient data, a QSAR approach could be
investigated to estimate the potency of the alkylated compounds. Such an
approach based on structure may be possible at this time e.g., Hansch 1991
(There are likely newer approaches available). However, if the potency of
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specific compounds is available the specific data should be used. To leave the
alkylated PAH out of the risk assessment completely would likely result in an
underestimate of risk. To include them using the parent compound potency is an
improvement but it is not known whether the risk would be under or over
estimated.
Charge 24: Please comment on the potential for human exposure to PAHs
through consumption of finfish and other seafood. The exposure of humans
to parent PAH through exposure to fin fish is expected to be low because of
known rates of biotransformation. However, humans might well be exposed to
the metabolites, which have not been studied for toxic effects. Such trophic
transfer of metabolites does occur in aquatic organisms (McElroy and Sisson
o
1989). While the impact is not known, the study did find that there was formation
of bound residues. This suggests that the metabolites are active for such modes
of action as carcinogenicity. There has been very little study of this issue. The
exposure to PAH from other seafood items, e.g., clams, mussels and lobster,
harvested from the HARS would be greater since these organisms have
generally lower ability to biotransform PAH than finfish.
Charge 25: What are the major sources of uncertainty associate with the
approaches described in section E? What alternative approaches would
reduce the uncertainties? How could these uncertainties be described and
accounted for in decision-making? While this is not my specific area of
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expertise, there are uncertainties associated with all of the factors used for the
calculation of risk. The values that are selected have both variability and
uncertainty associated with them. The impact of the variability for the parameters
is not specifically addressed in the section or in the calculation. What is created
is a deterministic solution. Several of the associated papers specifically attacked
one parameter or another in the models. The two most frequent concerns were
the decisions on the consumption rate and the duration of exposure. It seems to
me that EPA has examined each of these variables and selected values that
seemed appropriate to produce a protective (worst case) deterministic solution.
The rational for the selection of each of the values and the potential impact of the
variation was not readily spelled out in the document. It seems to me that for
regulatory purposes, the rational for the selection of specific values and the
impact of the known variability in those values should be discussed whether in
the main document or in an appendix. It would probably be good to do a formal
uncertainty analysis on the model and compare the results to the deterministic
solution to show that the deterministic solution is protective.
Charge 26: This is outside my area of expertise.
Charge 27: Please comment on the appropriateness of the proposed
approach for converting and using the analytical data for alkylated and
parent PAHs to estimate risk from all PAHs. It is expected that for a non-polar
narcotic mechanism of action that all the compounds should be summed. See
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the theoretical framework by DiToro and McGrath (2000). It is also clear that not
all compounds are equal when other mechanisms of action are considered, e.g.
photo induced toxicity or carcinogenicity. In this case, using the sum of PAH is
likely particularly necessary since the toxicity of the alkylated PAH have not been
well studied.
Charge 28: Do you believe that the "disaggregate" modeling discussed
above (and shown in Figure 4) for estimating human health HARS-Specific
Values for lead is appropriate? Would you recommend an alternative risk
assessment method be used given the information and data available? Do
you believe the method described has appropriately taken uncertainty into
account? Please elaborate. The approach of reducing the exposure to lead
through the trophic pathway to account for other sources of exposure to insure
that some safety level is not exceeded seems to be good practice. I am curious
why this is not being applied to all the toxicants that are being regulated.
Charge 29: This is outside my area of expertise.
Charge 30: Is the conceptual model for evaluating fish exposure to dredged
material at the HARS and human exposure through ingestion of seafood
appropriate and reasonable? How can the uncertainties associated with
the assumptions in this conceptual model be reduced? Please consider the
spatial and temporal elements of exposure in your discussion. The model
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for fish exposure is likely not the best that can be established today. Much more
use of regional data to establish the trophic links is important for improved
accuracy. The model selected for the trophic transfer calculation is a very
simplified food web and uses many parameters that have not been established
for the region. Further, the accuracy of the model and the uncertainty of the
model have not been clearly established. It would also be useful to perform
sensitivity analysis on the model to examine which factors are most critical and
insure their accuracy. The model structure should be examined to insure that all
the major processes are considered. It should be clear however, that while
spatial and temporal variation in exposure may well occur there is likely not
sufficient data to incorporate these variables. The model chosen should not
exceed the limits of the current data to support the model but the uncertainties
and potential limitations of the data and resulting model should be discussed.
The connection between the fish and humans is also a deterministic model and
there was significant discussion on the uncertainty of the factors that are a part of
that model. Both the trophic transfer and the risk assessment models should be
have rigorous uncertainty and sensitivity analyses performed to demonstrate that
the deterministic approaches do provide good protection.
References
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of organic chemical accumulation in aquatic food webs with sediment interaction.
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R., Campagna, P., and Mullin, J. 1999. Quantitative characterization of PAHs in
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PAHs-the 1994 Mobile Burn Study. Environ. Sci. Technol. 33:3100-3109.
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Hewitt, J.E. 1993. Bioaccumulation and elimination of chlordane by selected
intertidal benthic fauna. Environ. Toxicol. Chem. 12:733-742
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substituted and unsubstituted aromatic hydrocarbons of
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Comments Received from
Lynn McCarty, Ph.D.
March 25,2002
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L.S. McCarty Scientific Research & Consulting
94 Oakhaven Drive, Markham, Ontario, Canada L6C 1X8
Phone/Fax: 905/887-0772/66
1156868 Ontario Inc.
lsmccarty@rogers home.com
March 25,2002
Nancy Bonnevie
Peer Review Leader
Battelle, Duxbury Operations
397 Washington Street
Duxbury MA
USA 02332
Dear Ms. Bonnevie:
Please consider this letter my peer review of first phase of the project entitled "Proposed
Bioaccumulation Testing Evaluation Framework for Assessing the Suitability of Dredged
Material to be Placed at the Historic Area Remediation Site (HARS)." The first phase is focused
on human health effects. As requested I have prepared responses to the charge questions
provided.
Charge Questions
Overall Process
1. Throughout the proposed process, there are various uncertainties introduced. Please identify
the key areas of uncertainty that need to be addressed. Are the additional data sources or
parameters that could be used to address these areas? What methods are available for describing
and accounting for these uncertainties in the calculation of HARS-Specific Values? Of the
methods available, which would you recommend for consideration and why? Please consider the
implications of implementing theses methods in the regulatory framework. Please include an
evaluation of probabilistic and deterministic methods in your discussion.
One of my favourite quotes, and one that has particular pertinence to risk-based regulatory
activities such that under review herein, is one from the well-known statistician George Box:
"All Models Are Wrong But Some Are Useful"
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In the discussions below, it must be always kept in mind that, for the most part, the assumptions,
calculations, and models being employed have sufficient error, inadequacy, and uncertainty to
render them wrong in some absolute sense. Thus, it seems pointless to argue in great detail about
which assumption, calculation, or model is less wrong that another because at the end of the day
they are all still wrong. A more productive view is to determine which is more useful for the task
at hand for, after all, the main objective of a regulatory activity is not to be purely scientific, but
rather to employ scientific and other knowledge and considerations to achieve a policy objective.
The issue of the nature and sources of uncertainty, and the special case of variability in the
overall context of uncertainty, have been well reviewed by Finkel (1990), Rowe (1994), and
Hoffman and Hammonds (1994). The former also provides useful commentary on the role and
use of uncertainty in risk management decision-making. I recommend for inclusion in the
regulatory guidance a brief overview of the types and nature of uncertainty, the general approach
that will be employed to address each type in the guidance, and how these uncertainties are
addressed in the decision-making process. This latter point is particularly relevant to facilitate
risk communication to stakeholders for, as is noted by Finkel (1990), "... how one decides is
often more important than what one decides." This is in large part related to the fact that almost
invariably there is more unknown than known in the scientific component of risk-based
environmental decision-making.
It is clear that there are substantial knowledge gaps, both in terms of factual data and
fundamental concepts and theories, in the physical, chemical, biological, and toxicological
aspects needed to address the issues which are the subject of this regulatory activity.
Although more factual data are often helpful, especially where the existing data pool is meagre, a
more substantial and common barrier is deficiencies in theoretical aspects. For example, despite
the importance of mixture toxicity in the HARS evaluation scheme, no reasonably expected
increase in the amount of factual data can compensate for the fact that there is no broadly
applicable, generally-accepted theoretical framework for classifying modes of toxic action and
mechanisms of interaction (McCarty, accepted for publication). Thus, there is no sound scientific
basis for addressing mixture toxicity.
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For a variety of reasons I think the best approach in such cases is for any proposed guidance to be
as simple as is practically feasible and for the development and implementation to be open and
transparent to all parties. Also, where there is sufficient knowledge to move beyond the simplest
practical assumptions, calculations, and models, limit such moves strictly to the area where the
knowledge exists. Simple approaches with safety factors applied in proportion to perceived risk
do not obscure the lack of knowledge and does not give the process an unwarranted aura of
scientific respectability. Rather than limiting or degrading the input of science into regulatory
decision-making processes I believe that this keeps the contribution "honest" and helps to focus
scientists on the areas and issues most needful of additional scientific understanding. Although it
is my observation that the HARS process is reasonably successful in delivering on the open and
transparent aspect, it is struggling to keep things relatively simple. There is much research being
done in environmental toxicology and risk assessment and management methodologies and there
is the ongoing temptation to incorporate some of this work into regulatory guidance. However,
that which represents good research does not necessarily directly constitute good regulatory
guidance as the objectives are different.
My evaluation of probabilistic and deterministic methods is short. Although practitioners of
probabilistic methods always claim that their methods are better, more through, more scientific,
etc, as noted at the beginning of this section they are still wrong. In my opinion in many cases the
assumptions (i.e. default values) required to enable the use of sophisticated probabilistic models
make them less useful than simple deterministic methods employing "safety factors", since they
are neither simple nor readily communicable and thus fail the open and transparent test risk
communication test. Where probabilistic methods excel is in research, where they are critical to
determination input parameters and ranking of their influence on the outcome, I believe, at the
moment, probabilistic methods should rarely be an integral part of the regulatory guidance.
Regulatory guidance should typically employ relatively simple, typically deterministic, methods.
However, it is vital that these largely deterministic regulatory methods been thoroughly examined
and investigated via probabilistic methods and modified and/or subjected to limitations in scope
such that major sources of error identified by probabilistic studies are minimized to the extent
feasible. In short, probabilistic methods are vital to regulatory guidance, but their focus should be
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on optimizing relatively simple, relatively easy to communicate, largely deterministic, regulatory
guidance methods, rather than forming a substantial component of the actual guidance.
Proposed Additions to Analyte List: Alkylated PAHs
2. Is measurement of the 16 priority pollutant PAHs (i.e. parent PAHs) sufficient for
characterizing the risk associated with the total PAH bioaccumulated by organisms exposed to
dredged material proposed for placement at the HARS? Does the measurement of the alkylated
compounds significantly improve risk assessment of PAHs?
It is very difficult to answer either question in a general way. If the 16 priority PAHs represent
the bulk of the PAH contamination, and if any of the other PAHs present are of low potency
relative to the modes of action and response endpoints being considered for the priority PAHs,
then it may be possible that concentrations of priority PAHs provide a good basis for
characterizing the risk associated with PAHs. However, it is not possible to determine if the 16
priority PAHs, or even the 16 priority PAHs plus the 30 alkylated PAHs, are a good surrogate for
the total PAH risk without knowing more about the total PAHs in the area. The most basic
question in this regard is related to exposure and whether the ratio of the 16 priority PAHs to
total PAHs relatively constant and whether the composition of the non-priority PAHs relatively
consistent through time and space. Based on the review by Irwin et al. (1997, sections on PAHs
and alkyl PAHs) there are several important factors which must be considered. For example,
PAHs of petroleum origin have a greater proportion of alkyl PAHS to parent PAHs, unlike PAHs
originating from combustion, where parent PAHs predominate. Also, despite being products of
metabolic breakdown by organisms, alkylated PAHs are not necessarily of lesser importance as
they often have a greater lipophilicity (i.e. higher log Kow), greater environmental persistence,
and despite a general trend for metabolites to be less toxic than the parent compound, can be
more toxic than the respective parent PAH (e.g., methyl phenanthrene is more toxic than
phenanthrene).
The complexity of the problem is compounded by toxicological issues. The nature and degree of
effects are different depending on the target organisms and the effect endpoint being considered
(cancer versus non-cancer). The concern is direct non-cancer toxicity to aquatic organisms at the
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HARS while for human health the dominant issue is possible carcinogenic effects related to
consumption of contaminated aquatic organisms from the vicinity of the HARS. For this phase of
the review the focus is on the more difficult to interpret human health issue.
Given the poor level of knowledge on exposure, toxicity, and toxicity of complex PAH mixtures
for PAHs in general and alkyl PAHs in particular. I do not believe that it is possible provide any
credible answer to either question. To be more specific, the regulatory assessment of the 16
priority PAHs provides some measure of protection from exposure to specific PAHs but the
proportion of the total risk due to PAH exposure this represents in unknown. As well, addition of
the alkyl PAHs to the assessment process changes the risk estimate, but the accuracy and validity
of this process is also unknown. Again it is not possible to determine if this process would
improve the risk assessment of PAHs in any scientifically sound, quantitative manner.
3. Is the proposed adaptation of EPA Method 8270 (Appendix D) acceptable and appropriate for
regulatory decision-making? If not, what is an acceptable and appropriate method?
I am not an analytical chemist, so I do not feel qualified to make a judgement that is most
appropriately made by an professional chemist. However, as user of such analytical data I
approve of the Quality Control methods which appear in section 5.0 of the Appendix D
description. This procedure should provide good data forjudging the quality of the analytical
results for reported for PAHs. Although not mentioned here, I expect that the field sample
acquisition and handling/transport protocols which ensure that the samples taken are good and
fair, incorporates general measures that ensure that inadvertent or inappropriate contamination or
crosscontamination with alkyl PAHs does not occur. A particular concern is the production of
alkyl PAHs from parent PAHs by microbial action in the sample containers.
4. Under what specific conditions would the testing for alkylated PAHs for a particular project be
appropriate and warranted?
If general testing for alkylated PAHs is not implemented in the HARS protocol, project-specific
conditions for testing for alkylated PAHs should be subject to a policy directive on this matter.
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To determine site specific conditions some information is required to suspect that alkyl PAHs
may be present above a specified threshold where full testing is required. This could be based on
one or more the following decision criteria: levels of priority PAHS are above a specific
concentration (I expect that if priority PAHs are low, total PAHs, and alkyl PAHs are also likely
to be low), there is previous data to suggest sediment from this area or this type of sediment has
shown elevated alkyl PAHs levels, or limited screening testing for alkyl PAHs of sediments from
the site have shown levels above a specified threshold. Justification and validation of the process
and decision criteria would be needed, but it seems reasonable that if not all of the NY/NJ
harbour system is contaminated with high levels of alkyl PAHs, then not all of the sediments
need to be subjected to complete testing for alkyl PAHs.
5. What uncertainties would be introduced within the analysis for risk should alkylated PAHs be
included? What steps could be taken to account for these uncertainties in decision-making?
Given the likelihood the method using non-detects (as described in EPA/CENAN, 1997) will
result in an overestimate of risk, what are the implications?
Also see responses to question 2 and 10. A number of issues, both site-specific and general
would be introduced by inclusion of the recommended risk evaluation process for alkyl PAHs.
Perhaps the most important site-specific is a lack of knowledge about the fate and exposure for
these chemicals; specifically, are the concentrations of each alkyl PAH relatively constant in
time, space, and organism type in the vicinity of the HARS or do they vary in some predictable
manner, perhaps in association with the priority PAHs? General issues include changes in
lipophilicity and persistence between parent and alkyl form, variations in toxicity compared to
the parent, differences in toxicity mode of action, potency, and effect type to various receptors
(e.g., narcosis to some, cancer to others) and the lack of generally accepted mixture toxicity
framework. Of the two areas, the lack of toxicity knowledge is probably the most limiting and the
least likely to be affected by any additional HARS-specific information.
At the moment I feel that addition of the alkyl PAHs represents a effort that results in a false
sense of security and a unwarranted measure of scientific respectability since the toxicity-related
data and knowledge gaps are so large that the outcome of the process is largely dependent on the
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nature of the assumptions required to incorporate these PAHs in the assessment process. I
recommend that a small research program to determine the spatial and temporal relationship
between the priority PAHs and the alkyl PAHs and/or the total PAH loading in the NY/NJ
Harbour area be carried out. The objective is to establish if there is any reliable relationship
between these parameters such that the priority PAHs could be used as a surrogate for a larger
group of PAHs. If it is felt that there is an immediate need to do something, a "safety factor"
approach, say assuming that the total PAH risk is reasonably represented by a value which is the
sum of twice the risk for each of the 16 priority pollutants, may be a useful place to start and
represent the first iteration of a process of revision and refinement in which the experimental
work suggested would feed into. Also, continuing to follow research and development on PAH
TEQs, is recommended for if this reaches a reasonable level of development it would be useful in
extending the PAH guidance.
The issue of dealing with non-detects is addressed in question 10 but, generally speaking, it is not
scientifically appropriate to generate estimates of risk without clearly identifying factors
introduced to address various types of uncertainties. In this case a calculation method should be
employed which is estimated to give the most scientifically accurate estimate when non-detect
values are present in the data set and any "uncertainty" or "safety" factors should be incorporated
in a subsequent separate step where the policy direction for its use is clearly laid out.
Proposed Additions to Analyte List: Organotins
6. It is recognized that additional methods have been used for the analysis of organotins (e.g.,
Krone et al., 1989). Will the proposed analytical method (Rice et al., 1987) provide adequate data
of sufficient quality to assess relevant risks from organotins? If not, please provide
recommendations.
As indicated in my response to question 3,1 am not an analytical chemist, and do not feel
qualified to make professional judgements about the details of chemical analysis.
7. What special QA/QC procedures should be implemented to ensure the quality and usability of
the organotin data?
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As indicated above, I am not an analytical chemist, and do not feel qualified to make professional
judgements about the details of any special QA/QC procedures beyond recommending that the
standard sampling, transportation, storage, and process control and check values typically
employed to ensure data quality be used in with the organotins.
8. Under what specific conditions would the testing for organotins for a particular project be
appropriate and warranted?
As organotins have and continue to be used by marine vessels to reduce hull fouling it is virtually
certain that any active harbour with substantial vessel traffic will contain appreciable amounts of
these contaminants in sediments. Although watercraft in small recreational marinas may not be
significant sources of organotins, where, as is the case in the New York/New Jersey area, there is
heavy commercial traffic, sediments in the area will be generally contaminated. Thus, it would be
appropriate to test for organotins for all projects in the New York/New Jersey area. Although,
reductions in the use of organotins in hull fouling treatments are expected to occur over the next
decades, the extensive use and widespread contamination of harbour sediments around the world
means that this issue will need to be considered for a number of years.
The data presented in Appendix E of the Peer Review Package indicates that organotin levels in
the NY/NY harbour area are sufficiently elevated to cause molluscs to accumulate body residues
to a level where adverse effects have been observed in marine invertebrates. It is my view that
there is sufficient evidence to require organotin testing of all potential remediation material.
Proposed Addition to Analyte List: Coplanar PCB Congeners
9. If the approach for evaluating dioxin is modified, should it include the contribution for PCBs
with dioxin-like activity as proposed? If so how?
As part of the international discussion on dioxin-like toxicity, there is a consensus on the identity
and relative potency of co-planar PCB congeners that exhibit significant dioxin-like activity (van
den Berg et al., 1998). Although this consensus has and may continue to change as new data and
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understanding become available, it represents generally-accepted scientifically-based approach to
this issue. The U.S. EPA (U.S. EPA, 2001) addressed this issue and concluded that the toxicity
equivalence methodology is technically appropriate for evaluating risks to fish, birds, and
mammals, that the toxicity equivalence methodology reduces uncertainties and is less likely to
underestimate risks than are methods based on a single compound (e.g., 2,3,7,8-TCDD) or a
class of compounds (e.g., total PCBs), and that the uncertainties associated with using relative
potencies (RePs) or TEFs are not thought to be larger than other uncertainties within the risk
assessment process, but they should be better quantified. The planning committee concluded that
the results of the workshop support the use of the toxicity equivalence methodology in ecological
risk assessment. The committee also suggested the development of additional tools and data to
improve the methodology's implementation.
There a number of scientific unknowns and uncertainties associated with the application of the
currently formulated WHO-approved TEF approach, especially in using TEFs based on
biochemical and enzyme tests to extrapolate to effects and responses at the whole organism level
of organization. It is also known that, in addition to the orders of magnitude differences in TEFs
for some organisms (especially fish compared to mammals and birds) and some congeners, fish
appear to be particularly insensitive to mono-ortho PCBs (van den Burg et al., 1998). This lack of
validation and the considerable variance between TEFs for different species represents a
considerable concern to the use of this method in detailed regulatory risk assessments. However,
despite these concerns, it would be considered to be acceptable practice to move to this method
for evaluating dioxin-like toxicity using the most current WHO consensus on TEFs, including
TEFs for PCB congeners with significant dioxin-like activity, in the next revision of the EPA
Region 2 methodology for screening dioxin-like toxicity risks.
Comparison to Reference.
10. Please consider the policy for assigning values (at one half the detection limit) to tissue
residues that are reported as "
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reported as "
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effects or accumulation between the existing reference site and candidate sediments would be
useful as substantial differences in proportions of sand and silt/organic material can be associated
with differences in effects and/or accumulation, even when contaminant levels are not an issue.
Although the ideal would be to have the two reference sites in the testing protocol, I can
appreciate that this represents a significant increase in testing effort. Therefore, as an alternative
to routine use of the clean organically enriched sediment, a detailed experimental examination of
this issue employing the three sediment types and all of the typically monitored chemicals, could
provide knowledge about the influence of the various sediment types on organism accumulation,
and may be the basis for allowing the current clean sand control only methodology to continue.
10.2. The question concerning the use of detection limit and/or half-detection limit in statistical
analysis.
This question is readily resolved. The practice of using arbitrary cutoffs such as the detection
limit or half of the detection limit has been roundly criticized by statisticians as being
inappropriate and misleading. Although I have a number of papers on this issue, the two noted
herein provide review, discussion, and recommendations. Newman et al. (1989) compared eight
methods for dealing with censored data using 3 sample data sets. Two approaches, the maximum
likelihood estimate (MLE) with bias correction and regression order statistics (ROS) performed
much better than substitution or deletion methods. Helsel (1990) also offers a good examination
of the issue and recommends MLE methods. Therefore, the policy of assigning arbitrary values
for nondetects in any analysis, tissue or otherwise, should be stopped and replaced with an
appropriate, statistically sound method. The result should be a substantial reduction in estimation
errors where the previous policy had been employed.
11. Is the use of functional groupings in statistical comparisons to reference appropriate and/or
preferable to statistical comparisons using individual contaminants for the purposed of risk
analysis?
It is not possible to answer this question in the general case, for, to do so accurately requires a
clear definition of "functional groupings" for in the case in question. For example, if the
grouping is based on is consistent with the toxicological endpoint which is the basis or reason for
doing the comparison, the chemicals in the grouping are thought to have the same mode of toxic
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action at a similar dose level range and the toxicological endpoint under study, and grouping is
on the basis of molarity not mass (toxicity is largely a function of the number of molecules not
their weight), then statistical analysis of the group is on a reasonably sound basis. The grouping
of chemicals that cause dioxin-like toxicity and the use of the TEF potency adjustment is an
example of a generally accepted functional grouping. On the other hand the PAHs in general are
not a functional grouping as they have subgroups based on both chemical structure similarities
and toxicological aspects; specifically, subgroups defined according to regulatory (priority) and
chemical structure basis (alkyl PAHs) and according to carcinogenic or noncarcinogenic effect
endpoints. Also, the proposed TEF potency scheme based on benzo[a]pyrene is not a generally
accepted methodology.
Although the use of functional groupings in regulatory evaluations can be useful, the rationale
must be clearly indicated in each case for as noted in the response to question 1, they represent a
modelling approach and are surely wrong at some level of sophistication. There may be an
external source such as the WHO TEF scheme for dioxin-like toxicity or a specific grouping may
be identified for a particular project, such as the benzo[a]pyrene potency scheme identified in the
HARS material. However, be the nature and justification for each functional grouping scheme
must be explicitly provided. Given the nature of the current toxicological limitations in
regulatory risk assessment the basis for functional groupings is a least as much driven by
administrative convenience as it is justified by scientific knowledge.
As an additional note, although there is and appropriate discussion about the significance of mass
versus molar units in toxicology (see Appendix B, pg 53), I believe that all manipulations should
be carried out in molar units and conversion to mass units should be left to the very last. Working
with converted or adjusted mass units, although keeping the spirit of the importance of molar
units generates even more confusion in what is a confusing aspect of toxicology for many people.
Also, it is important to ensure that TEFs are applied to data in the appropriate units, as TEFs for
mass units are different than those designed to be used with molar units and TEFs based on
exposure-based doses are different than those for received doses such as whole-body organism or
tissue concentrations.
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Adjustment to Steady State: Organic Compounds
12. Is it appropriate to apply a multiplier based on log Kow for these compounds (organics), or
are there other specific data that can be used to estimate steady state? If so, please identify.
The magnitude of bioaccumulation and time taken to achieve steady-state in organisms exposed
to a chemical are affected by a number of biological factors such as body size, lipid content,
temperature, and chemical-specific metabolic degradation rate as well as the log Kow of the
chemical. Metabolic breakdown in particular can be highly variable between species and
chemicals. There are empirical data for some of the organic chemicals being considered so that
there is some information to indicate that PCBs and the pesticides on the list of organic
chemicals are relatively recalcitrant to biodegradation. In fact, they are part of the database that
generated the log Kow-bioaccumulation relationships being employed. However, there is not
much in the way of reliable observed data for the priority PAHs and, especially the alkyl PAHs.
As the PAHs are the most likely organic chemicals on the list to be subject to appreciable
metabolic degradation by organisms, the use of a log Kow-based estimation processes for them
may be less reliable than for the others. However experimental data from the New York Bight
area looking at bioaccumulation of 15 PAHs from sediment into clams reported reasonable
predictability in bioaccumulation (McFarland, 1998). Despite any uncertainty in the prediction
and adjustment process, the estimates produced for PAHs by a log Kow-based procedure are not
likely be greater than that which actually occurs, as metabolism will reduce the magnitude of the
steady-state relationship to some value below that predicted. Thus, this estimation process will
not likely contribute to an underestimation of risk.
13. Given the increased hydrophobicity of alkylated PAHs, is the use of the correction factor
associated with the corresponding parent an appropriate approach for estimating steady state
residues of alkylated PAHs? if not please elaborate.
See also the response to questions 2, 5,12,23,24, and 27. Increasing the hydrophobicity will
alter the bioaccumulation kinetics proportionally. However, the steady-state correction factor is
only a rough approximation in the first place and other factors, especially organism-specific
differences in accumulation efficiency and metabolic degradation will also effect estimates. If
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this adjustment process is used then probably acceptable as a first approximation for screening
purposes, when there is no indication of substantive differences in environment fate, absorption
and metabolism compared to the parent Kow-based correction factor. Given the uncertainties and
lack of knowledge about the toxicity of the alkyl PAHs it is likely that assumptions about
changes in hydrophobicity due to alkylation will be only a minor influence in the risk
determination process.
14. For the DDT derivatives and dieldrin, please comment on the appropriateness of using M.
nasuta data rather than N. virens-specific data in the estimation of steady state multipliers.
Question is whether bioaccumulation data on DDT and dieldrin for a bivalve mollusc (.Macoma
nasuta) can be appropriately employed as a surrogate for bioaccumulation in an polychaete worm
(Neris virens). The main determinants of bioaccumulation of neutral organic chemicals such as
these is exposure route, bioavailability, assimilation efficiency, lipid content, and
metabolic degradation. Although it would be possible to make a specific evaluation of these
species and these chemical using the above parameters, there is little need to attempt such
precision when there is substantial known variability within the general relationship.
Tracey and Hansen (1996) using published information from three laboratory and five field
studies containing 27 vertebrate and invertebrate species (including the two in this question) and
4054 BSAF values (including pesticides) evaluated BSAF values for various species both within
and among habitat groups, and found that the lipid-normalized BASF values were generally
similar for all species within but not between the major chemical groups of PAHs, PCBs and
pesticides. For the pesticide grouping the median BASF was 2.58 and 1.62 M. nasuta and N.
virens, respectively. Although the general similarity is for the magnitude of the steady-state
relationship, not the kinetics of achieving it, this report does demonstrate considerable similarity
in bioaccumulation for these two species and supports the use of substituting one species for the
other in this bioaccumulation screening process.
15. Are the approaches taken to adjust organic contaminant bioaccumulation data to steady state
adequate? Do the proposed multipliers agree with previously published studies (i.e., do they
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appear reasonable)? If not, please elaborate.
Also see the response to questions 12 and 13. The approaches taken to adjust steady-state
estimates based on long-term exposures to 28-day exposures reflect the current understanding of
relationship between log Kow and bioaccumulation for many persistent neutral organic
chemicals as well as the influence of modifying factors, including as bioavailability, absorption
efficiency, lipid content, and biodegradation, that can affect both the magnitude of
bioaccumulation and the time taken to reach steady state. Although no general rule can address
all of the influences acting on a particular chemical and/or organism, and some differences
between predicted and measured values have been observed for some chemical-organism
combinations when experimental work has been carried out, the proposed multipliers appear to
be within the range of generally accepted values. However, as noted elsewhere the use of this
adjustment process for alkyl PAHs is more uncertain and error-prone as there is little
experimental data to confirm if and when they follow the same relationship.
16. What are the major sources of uncertainty associated with the approaches? What alternative
approaches would reduce the uncertainties? How could these uncertainties be described and
accounted for in decision-making?
Also see the responses to questions 12,13,14, and 15.The major sources of uncertainty are the
variability in knowledge among various chemicals of concern and a general lack of knowledge of
the complex environmental fate, distribution, bioavailability, accumulation, internal distribution,
and metabolic breakdown for these chemicals in the living organisms of interest. Other than the
development of more data and improved scientific understanding of these issues, I see no
alternative approach which will substantially reduce the uncertainties. The use of conservative
evaluation and modelling procedures in both risk assessment and management, as is generally
employed in the processes outlined for use at the HARS, is currently the only feasible approach
currently available for addressing uncertainty. As noted elsewhere, the uncertainty associated
with steady-state bioaccumulation is likely low relative to other aspects of the exposure
estimation process and the toxicity assessment process.
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17. In your opinion, is the methodology followed to derive a steady state multiplier for non-
essential metals (i.e., a factor of three) scientifically appropriate (Appendix G)? Please elaborate.
Do you have any recommendations of additional or alternate methodologies or
information than can be used either to supplement or replace the proposed method?
This appears to be a logical, empirically-based method to justify a steady-state multiplier of 3 for
nonessential metals in the New York Bight area. Evaluation of available metals bioaccumulation
data, and generation of new data where deficiencies exist, may strengthen the confidence in this
approach, refine the multipliers, and address the issue that the field study on which this method is
based did not establish the actual exposure durations for sediment-organism samples employed.
However, given the larger uncertainties elsewhere in the assessment process, this uncertainty is
likely small and is not deserving of any priority.
Human Health Evaluations: Overall
18. Please comment on each factor listed above (and in Table 5) as to its appropriateness for use
in the equations listed above. Would you recommend additional factors? Would you change or
modify the equations as written above? If so, how?
Cancer Potency Factor/RfD: This is a required toxicological factor for risk estimation for cancer
and noncancer endpoints, respectively. The use of current IRIS values and the proposed lead
factor appear appropriate. As noted in the response to questions 2, 5, and 23,1 do not have
confidence in the approach to alkyl PAH toxicity.
Seafood Consumption: Although making such determinations is a daunting task, the estimate 7.2
g for daily consumption of fish from the vicinity of the HARS has been reasonably justified.
However, the question raised by the NYSDEC indicating that they felt at the value should be
11.3 g per day, should be resolved (See the response to question 30).
Exposure Duration: An exposure duration 70 years is a very conservative one as it is unlikely
than any individual will ever achieve a 7.2 g daily consumption of fish from the HARS site for
essentially an entire lifetime. A lesser, more plausible value should be considered such as the
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EPA Superfund values of 350 days per year for 30 years.
Site Use Factor
The suggestions made elsewhere (see the response to question 30) regarding improving this
estimate by incorporating consideration of fish habitat/foraging areas have merit and should be
considered to improve this factor in the equation.
Whole-Body to Fillet Factor
See also the answer to question 20.1 believe that the use of this correction factor has not been
sufficiently justified and should not be employed in the HARS risk estimation process.
Trophic Transfer Factor
Trophic transfer factors are a work in progress since both the theory and concepts are still being
developed and refined and new data is still being generated. The outline of the development of
these factors presented in Appendix I indicates that a good examination and consideration of the
state of the science was made and reasonable values for factors chosen.
Exposure Equations
See also the answer to question 30. The equations appear to incorporate the important factors
identified in an adequate manner to generate acceptable deterministic risk estimates.
Overall, I don't see a pressing need to recommend any additional factors. As for modifying the
risk equations, there are many ways to achieve cancer and non-cancer risk estimates, but these
appear to adequately address the major factors in a deterministic manner. That being said, the
concerns raised in the submission on behalf of the Nation's Port should be addressed. Although I
think that several of the points raised in item for regarding the risk equations are related to
alterations/simplifications related to HARS-specific consideration, the points about use of
nonstandard nomenclature and lack of consideration of EPA Data Quality Objectives in selecting
model parameters must be rebutted or the equations/parameter assumptions revised.
19. Are the methods used to derive the human health exposure parameters and assigned values
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discussed in Section E appropriate (please review the referenced appendices)? If not please
elaborate. How should these factors be factored into the risk analyses and decision-making?
This question is answered elsewhere, see the responses to questions 18, 20 through 30.
20. Is the approach taken to relate fish whole body and fillet concentrations scientifically
appropriate? If not, what method would you recommend?
The approach is based on empirical relationships which appear to be generally reasonable, but I
do see a particular problem with metals. The human health risk from arsenic and chromium is
largely due to inorganic forms where the whole-body to fillet adjustment suggested can be
logically justified. However, the human health risk is not due to metallic forms of mercury but to
organic forms such as methylmercury. As methylmercury behaves more as a lipophilic organic
compound than an inorganic compound, the rationale should be different. The Bevelhimer et al.
(1997) reported cited in the HARS document provides data that shows that the concentration of
total mercury in the fillet is greater than total mercury levels in the whole fish. This is the basis of
the 0.7 adjustment factor. For many fish I would expect that exactly the opposite would be true
for organic mercury as removing the skin and guts would remove much of the depot fat in the
fish. Since the fillet is now much leaner than the whole fish, the concentration of methylmercury
should be lower. This relationship is complicated by seasonal variation in lipid levels in fish and
differing locations for storing lipids. For example, some fish such as salmonids, store much of
their fat between muscle groups and this would not be removed by filleting. Additional
confounding may occur if inorganic mercury is converted by human gut bacteria into organic
forms to any significant degree. Unless more information and analysis is carried out I suggest that
no adjustment be carried out for mercury.
In general, although the rationale for adjusting for whole-fish to fillet is logical, the amount of
influence this adjustment has is modest and subject to variations that are likely larger than the
amount of correction. As well, the assumption that all fish are filleted before consumption is not
likely true. For these reasons I believe that this correction factor should not be employed in the
HARS-specific risk estimation process.
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21. Could the analysis be improved by focusing on a smaller number of key fish (seafood)
species at the HARS? What characteristics should be used to select these key species?
The estimation of fish consumption from recreationally-caught fish would probably best be
improved by consideration of the foraging/habitat of the important recreationally caught fish in
the HARS area.
22. In your opinion, is the approach for assuming total metal to be the most toxic form
appropriate and reasonable? Should metal speciation/complexation be considered in the
assessment of metals bioaccumulation, trophic transfer, and human health risks? Is the proposed
approach for evaluating methyl mercury appropriate? Are there alternative analytical or risk
assessment techniques available that would improve the risk assessment of metals? Is the
multiplier proposed for adjusting measured concentrations of arsenic appropriate and reasonable?
The use of total metal as the most toxic form is not an acceptable generalization as, in many
cases, it is a fraction of the total that is in a specific form that is the primary toxic agent; e.g.
mercury and chromium. The form (species/complex) of many metals affects toxicity by either
affecting bioavailability to organisms and/or toxic potency or mode of action. If metal speciation
is ignored a significant contribution to uncertainty may be introduced into the assessment process
and it is likely that the risk assessment will be overly cautious. It may be possible to use total
metal concentration as the regulatory metric in some cases but this should be justified on a metal
by metal basis.
However, the above being said, in the interests of efficiency and economy a tiered approach for
each metal, such as proposed by EPA Region 2, is an acceptable methodology. My major concern
here is the way this is structured. Initially, the total metal residue is used in the risk assessment
calculations and, if this estimate indicates excessive risk, the opportunity exists for a proponent
to generate metal species/complex data which should provide a more realistic risk estimate as
part of a final judgement process. The problem is that important details of the more detailed,
enhanced process are not specified in advance. For example, although some forms may not be as
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toxic as others, there still may be some contribution and potency factors for various forms should
be specified in advance. As well, exposure/bioavailability factors may vary between forms and
some estimates should be provided for the other forms. If a tired approach is to be proposed,
details of both the simple and the detailed process must be provided in advance so that despite
whichever method is employed, data requirements are clear, the evaluation process is laid out,
and judgements can be made promptly.
The use of a 0.1 factor to adjust total accumulated arsenic in organisms to reflect the
toxicologically important component of inorganic arsenic in the total, and the subsequent use of
specific potency factors for inorganic arsenic in the risk assessment process with that adjusted
estimate is a good example of how the evaluation process for other metals might be improved by
addressing the most toxic form rather than using the total concentration. As noted elsewhere (see
response to question 30), the use of a 0.1 factor for arsenic appears to be overly conservative and
a value in the range of 0.01 to 0.04 appears to be more appropriate.
23. Is the assumption that the potency of alkylated PAHs can be estimated by the potency of the
parent PAH appropriate? Is this assumption likely to result in an under- or overestimate of the
risk associated with the alkylated PAHs?
See also the responses to questions 2 and 5. As a general rule it is not possible to make the
assumption that the potency of alkylated PAHs has some constant relationship with the
respective parent PAHs. Some alkyl PAHs are more or less potent for the same mode of action
and some may also cause toxicity by a different mode of toxic action than the parent compound.
Thus, I am not aware of a reliable method for predicting alkyl PAH toxicity from the parent
compound and both over- and underestimates of risk are likely to occur in the group. However,
perhaps there is a method which may improve this approach by providing some evaluation and
reducing reliance on pure assumption. Specifically, an evaluation of the various types of adverse
effects elicited in the exposed organisms at various dose levels for the parent and respective alkyl
PAHs may provide sufficient information to either confirm that the assumption is approximately
correct, determine a general rule that there is some relatively constant proportionality that can be
employed, or provide the basis for a simple TEF scheme.
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24. Please comment on the potential for human exposure to PAHs through the consumption of
finfish and other seafood.
See also the response to question 18. As most PAHs are readily metabolized by organisms, much
of the PAH content in seafood organisms is in associated with the gut, especially bile, and
typically is not available for human consumption when the organisms are cleaned. Where
seafood, such as shellfish, are eaten whole the cooking process should reduce the PAH content.
Although some types of cooking may substantially decrease PAH content, certain styles of
cooking ( BBQ, frying) also create copious quantities of PAHs. Thus, there is the potential for
human PAH exposure from other pathways and sources to largely overshadow the contribution
from consumption of seafood from the HARS area. Perhaps the best way to determine the
significance of consumption of seafood from the HARS area is to make an estimate of the
proportion of the overall PAH exposure that target human consumers obtain from HARS
seafood. If this is relatively small then the effect of employing simple assumptions in the risk
assessment process may be minor.
25. What are the major sources of uncertainty associated with the approaches described in
Section E? What alternative approaches would reduce these uncertainties? How could these
uncertainties be describe and accounted for in decision-making?
This question is rather open-ended and difficult to answer in a general sense. As noted in the
response to question 1, when dealing with models there is no "right" answer and the uncertainties
vary for the various chemicals being considered so general recommendations for uncertainty
reductions are inappropriate. Specific responses to various areas of uncertainty are provided in
response to specific charge questions.
26. What is your recommendation for evaluating the potential toxicity of organotins? Should they
be evaluated as individual compounds? Summed as total? Should there be some consideration of
relative toxicity?
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It appears that adverse effects on marine invertebrates such as shellfish are more sensitive to
organotins and effects would be expected to occur a lower levels than those producing adverse
effects in mammals. Thus, measures to reduce effects in marine organisms would likely provide
more than adequate protection to humans from the aquatic exposure pathway. I believe that
organotins should be assessed using a potency adjusted sum of the contribution of the congeners
examined based on a response endpoint for adverse effects on the most sensitive organism.
27. Please comment on the appropriateness of the proposed approach for converting and using
the analytical data for alkylated and parent PAHs to estimate risk from all PAHs.
This question has been addressed by the responses to questions 2,5, and 23. However, to
review, the proposed approach is to attribute to the indicated alkyl PAHs the environmental
fate/distribution of their respective parent PAHs and a toxicity potency either at some fraction of
that of benzo(a)pyrene for suspected carcinogens or that of their respective parent PAH for
noncarcinogens so as to facilitate risk assessment. The short answer is that this is an assumption
that does not appear to be supported by existing knowledge. Specifically, some alkyl PAHs are
known to be more potent and/or cause toxicity by a different mode of toxic action, alkylation
changes the hydrophobicity to different values than the parent PAHs, and susceptibility to
metabolic degradation is often changed by alkylation. Finally, there are many more PAHs than
those indicated here and there is no clear justification provided to indicate that these are either
representative of or an adequate surrogate for all PAHs.
Thus, despite the attractiveness of the proposed approach from a policy point of view, it must be
considered to be poorly supported as a scientifically-based approach for dealing with the risk
from all PAHs. Given this, there is some question of the utility of obtaining detailed accurate
PAH chemical analyses for additional PAH congeners when the fate and toxicity knowledge is so
poor as to make the risk estimates little more than guesses with large uncertainty associated
them. I believe it would be best to proceed with the 16 priority PAHs as an identified PAH risk
which will be evaluated by the cancer and noncancer approaches indicated and forgo
consideration of risk from additional alkyl PAHs and the total PAH risk until a more sound
knowledge base exists. If there remains a strong desire to address the total PAH issue then it
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would be appropriate policy to assign an arbitrary safety an/or uncertainty factor to provide some
additional measure of protection which addresses the issue that only a fraction of the total PAHs
are being examined. The safety/uncertainty factor approach clearly communicates the scientific
limitations in the process rather than the proposed methods which tend to obscure the problem.
Human Health Evaluation: Comparisons to HARS-Specific Values
28. Do you believe that the "disaggregate" modelling discussed above (and shown in Figure 4)
for estimating human health HARS-Specific Values for lead is appropriate? Would you
recommend an alternative risk assessment method be used given the information and data
available? Do your believe the method described has appropriately taken uncertainty into
account? Please elaborate.
The "disaggregate" model employed for lead is appropriate for chemicals where no RfD or
cancer potency factor is available. Adequate knowledge exists in several areas; specifically, fate,
exposure, and toxicology and there is a sufficiently well-developed method to develop what is
essentially a site-specific human health guideline for lead. Although there is some debate about
some of the assumptions used in the analysis, if there is any revision to any of theses factors, the
change can be readily incorporated and a revised lead value generate. As noted in the response to
question 1 it can safely be assumed that this model is wrong, but various revisions or alternatives
to it would also be wrong. However, the process draws on available knowledge and information,
processes it in a relatively simple manner, and presents it in a easily understandable way. Thus, it
must be judged to be useful and one of several possible appropriate methods for achieving
protection of human health from exposure to lead from the HARS site.
Human Health Evaluations: Consideration of Combined Effects
29. In your opinion, are the methodologies and equations described above appropriate for
estimating total carcinogenicity and combined non-cancer impacts of contaminant mixtures
accumulated from dredged materials proposed for use as Remediation Material at the HARS?
Although there is some consideration of mixture toxicity for some groups (dioxin-like toxicity,
PAH toxicity), there is no explicit consideration combined effects of all of the substances being
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considered. In essence the issue of combined effects and possible mixture interactions has not
been dealt with. However, as there is no generally accepted, broadly applicable method to do this
it is not surprising that current method has been employed. In short, this assumes that if the risk
from each substance and/or toxicologically-defined group of substances and each type of toxic
action and adverse effect is at a level that is generally considered to be acceptably low for the
most sensitive and/or most important receptor, then the combined effect for all of substances
collectively is also assumed to be low. Until further developments in toxicological knowledge in
this area occur, this type of approach is the state of the art in consideration of combined effects.
30. Is the conceptual model for evaluating fish exposure to dredged material at the HARS and
human exposure through ingestion for seafood appropriate and reasonable? How can the
uncertainties associated with the assumptions in this conceptual model be reduced? Please
consider the spatial and temporal elements or exposure in your model.
Aspects of this question have been addressed in several of the responses to previous questions so
I decided to use this as an opportunity to review material included in the Supplemental Section of
the HARS Peer Review Charge Document. These submissions made on behalf of several
agencies and organizations will be reviewed and important points addressed as a basis for
answering this question. Although some of these responses may be addressed to a specific
chemical of concern (e.g., PCBs), usually the issues raised have implications for the assumptions
and processes used and therefore have general applicability.
1. U.S. Army Corps of Engineers.
The first paper is an examination of probabilistic foodchain exposure and risk model. It is a
useful examination of separation of variability and uncertainty and the utility of ranking
parameters for variability/uncertainty so as to direct data collection to reduce overall uncertainty
in the models and in the decision-making process. The second paper considers the influence of
spatial and temporal patterns in exposure modelling. It also is a useful examination investigating
the influence of fish foraging/habitat area and management site characteristics on fish
consumption models.
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My personal opinion is that the paper on variability and uncertainty provides helpful guidance for
future data collection and refinement of risk estimation processes. The argument for the inclusion
of consideration of fish foraging/habitat area and management site characteristics on fish
consumption models is persuasive. Such revisions should be considered for inclusion in the
exposure models used for HARS.
2. Port Authority
Two useful papers similar in scope and content to those submitted in the U.S. ACE section were
included here. See above for the evaluation.
3. Nation's Port
Several fact sheets were submitted.
Fact Sheet 1. Deficiency 1, bioaccumulation of PCBs and other hydrophobic organics by fish will
be primarily from food with direct absorption by water representing a modest to negligible
contribution. Deficiency 2, as noted elsewhere, consideration of fish foraging/habitat in the
assessment is useful and should be incorporated into the HARS methodology. Deficiency 3, the
worms are considered representative of organism in fish diet and the recommended division into
worm and nonworm species is not appropriate in the current methodology. It would be useful
where more detailed foodchain relationships were used in the model. Deficiency 4,1 believe that
the proposed HARS equations are adequate but would be interested to see a rebuttal by the U.S.
EPA. Deficiency 5, the model parameter issues are addressed in other responses, but I would be
interested to see a rebuttal by the U.S. EPA on the Data Quality Objectives point.
The alternative equations, although somewhat different that recommended for HARS use, do not
appear to be dramatically different, although uptake directly from water is included. Without a
detailed examination I cannot be sure that they will generate substantially different outcomes.
Fact Sheet 2. The issue about improved fish foraging/habitat methods in the assessment is useful
and, as noted elsewhere, should be incorporated into the HARS methodology.
Fact Sheet 3. Setting of the maximum allowable dioxin level in worms is a policy issue that
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extends beyond the scope of the HARS process and should be dealt with elsewhere.
Fact Sheet 7. The issue of the use of fish populations versus individual fish is a matter of policy,
but I feel that a population basis is likely to provide a more realistic basis for decision-making.
Fact Sheet 9. The paper cited by Feijtel et al. (1997) is a newer reference that provides a useful
review and update of the bioaccumulation issue. My views on this issue are noted in my response
to questions 12,13, and 15.
4. NYSDEC
Point 1. The discrepancy between the fish consumption estimate applicable to HARS - 7.2 g per
day versus 11.3 g per day - must be resolved and the NYSDEC argument is persuasive.
Point 2. No comment
Point 3. No comment.
Comments on TEF. See my response to question 9.
5. Clean Ocean Action
Fourteen recommendations were made. I have prepared a very brief response to each.
1. Inclusion of coplanar PCBs. Addressed elsewhere in my responses.
2. Inclusion of alkyl PAHs and organotins. Addressed elsewhere in my responses.
3. Total mercury analysis. Addressed elsewhere in my responses, but I must note that
methylmercury is the toxic agent.
4. Conservative steady-state adjustment. I believe that parameters should be estimated as best as
possible by the current science and that the place for safety factors is after the initial assessment
so that the safety factor component is explicit and safety factors embedded in various parameters
are less likely to be compounded inadvertently.
5. Trophic transfer factors. No comment to this general comment.
6. Gut contents. The debate about this continues as the influence varies for chemicals and
species.
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7. Cumulative exposure. The point that exposures should consider the total of background
exposure plus a HARS-specific contribution is a valid one.
8. Site use factor. Addressed elsewhere in my responses.
9. The RME population of recreational fishermen has been selected. Some consideration of the
recreationally-caught fish consumption by women and children in these families would be
appropriate.
10. The use of a particular cancer risk level is policy not science and not appropriate for me to
comment on.
11. Target seafood. I believe that appropriate recreationally caught fish are being considered. I
have not seen information to justify inclusion of HARS shellfish into the exposure of the
recreational fishermen who represent the RME group.
12. Lobster. I have not seen information to justify inclusion of HARS lobster into the exposure of
the recreational fishermen who represent the RME group.
13. Target species characteristics. I believe this has already been done.
14. Validation of models is an oneoine general issue and will not be resolved at HARS.
6. NJDEP
Nine issues were commented on. I have prepared a very brief response to each.
1. The selection of a cancer risk level is policy not science.
2. Fish consumption rate. Addressed elsewhere in my responses.
3. Site use factor. Addressed elsewhere in my responses.
4. Whole body to fillet factor. Addressed elsewhere in my responses.
5. Mercury trophic transfer factor. Addressed elsewhere in my responses.
6. Inclusion of coplanar PCBs. Addressed elsewhere in my responses.
7. PAH and PCB toxicity through narcosis. This appears to be a question directed at the
ecological risk assessment which the subject of a future review process.
8. HARS specific ecological values. This appears to be a question directed at the ecological risk
assessment which the subject of a future review process.
9. Combined effects. No comment.
7. SEA
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The submission addresses some important topics in are relatively general manner. These issues
are addressed in more detail elsewhere.
8. U.S. ACE
The paper provides a useful examination of the effect of the proportion of arsenic in fish tissue
which is assumed to be in the most toxic inorganic form. The paper makes a good argument that
the assumption that !0% of the total fish tissue concentration of arsenic is in an inorganic form is
high and that the scientific literature available suggests that 1% or much more conservatively 4%,
would be a more appropriate value.
9. U.S. ACE SOWs
Three proposed additional data collections projects. Each of these would provide useful
additional data but the decision to carry out this work is not primarily a scientific decision.
Literature Referenced
Finkel, A.M., 1990. Confronting Uncertainty in Risk Management: A Guide for
Decision-Makers. Center for Risk Management, Resources for the Future, Washington DC.
Gardiner, W.W., E.S. Barrows, and J.Q. Word, 1996. Ecological Evaluation of Proposed
Reference Sites in the New York Bight, Great South Bay, and Ambrose Light, New York.
Battelle Memorial Institute. U.S. DOE Contract DE-AC06-76RLO 1830.
Helsel, D.R., 1990. Less than obvious statistical treatment of data below the detection limit.
Environ. Sci. Technol. 24:1766-1774.
Hoffman, F.O. and J.S. Hammonds, 1994. Propagation of uncertainty in risk assessments: the
need to distinguish between uncertainty due to lack of knowledge and uncertainty due to
variability. Risk Anal. 14:707-712.
Irwin, R.J., M. VanMouwerik, L. Stevens, M.D. Seese, and W. Basham, 1998. Environmental
Contaminants Encyclopedia. Electronic Document. National Park Service, Water Resources
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Division, Fort Collins, Colorado.
McCarty, LS, accepted for publication. Issues at the Interface Between Ecology and Toxicology.
Toxicology.
McFarland, V.A., 1998. Estimating Bioaccumulation Potential in Dredged Sediment Regulation.
In: U.S. EPA, Proceedings National Sediment Bioaccumulation Conference, September 1996.
EPA 823-R-98-002. U.S. Environmental Protection Agency, Washington DC. pp. 2(7)-2(18).
Newman, M.C., P.M. Dixon, B.B. Looney, and J.E. Pinder HI, 1989. Estimating mean and
variance for environmental samples with below detection limit observations. Water Res. Bull.
25:905-916.
Rowe, W.D., 1994. Understanding uncertainty. Risk Anal. 14:473-750.
Tracey, G.A. and D J. Hansen, 1996. Use of biota-sediment accumulation factors to assess
similarity of nonionic organic chemical exposure to benthically-coupled organisms of differing
trophic mode. Arch. Environ. Toxicol. Chem. 30:467-475.
U.S. EPA, 2001. Workshop Report on the Application of 2,3,7,8-TCDD Toxicity Equivalence
Factors to Fish and Wildlife. Risk Assessment Forum, U.S. Environmental Protection Agency,
Washington DC.
Van den Berg, M., L. Birnbaum, A. Bosveld, B. Brunstrom, P. Cook, M. Feeley, J. Giesy, A.
Hanberg, R. Hasegawa, S. Kennedy, T. Kubiak, J. Larsen, F. van Leeuwen, A. Djien Liem, C.
Nolt, R. Peterson, L. Poellinger, S. Safe, D. Schrenk, D. Tillitt, M. Tysklind, M. Younes, F.
Waern, and T. Zacharewski, 1998. Toxic Equivalency Factors (TEFs) for PCBs, PCDDs, PCDFs
for Humans and Wildlife. Environ. Health Perspect. 106:775-792.
Yours sincerely,
L.S. McCarty, Ph.D.
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Comments Received from
Anne McElroy, Ph.D.
March 26,2002
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Overall Process
1. Throughout the proposed process, there are various uncertainties introduced. Please
identify the key areas of uncertainty that need to be addressed. Are there additional data
sources or parameters that could be used to address these areas? What methods are
available for describing and accounting for these uncertainties in the calculation of
HARS-Specific Values? Of the methods available, which would you recommendfor
consideration and why? Please consider the implications of implementing these methods
in the regulatory framework. Please include an evaluation ofprobabilistic and
deterministic methods in your discussion.
This question should really be posed last. I have chosen to answer it after addressing the
other questions below. I am not an expert in risk assessment, and have virtually no
experience with human health risk assessment. My comments are primarily based on my
experience in determining bioavailability, metabolism, and toxicity of organic
contaminants to marine organisms. As discussed several times below in response to
questions concerning the uncertainty in the data used in the risk assessment, without
including variability in the model parameters into the model, I don't think uncertainty can
be adequately addressed. Furthermore, without some sort of sensitivity analysis, how can
you determine which parameters matter most to the risk assessment. Merely adding in a
safety factor of 10 may not adequately adjust for some of the variability you would
expect to see in some of these parameters.
One of my concerns in this whole process is the use of a sandy area as the reference site.
I have reviewed the Battelle report prepared by Gardiner et al. (1996) concerning
evaluation of reference site for use in the HARS. Although they showed no significant
toxicity in amphipod tests conducted on sandy sediments, I still do not think it is
appropriate to use a sandy sediment as a reference for either toxicity or bioaccumulation
from what are primarily fine (1-2% organic carbon) sediments. Although it would
appear that comparing bioaccumulation from a muddy test sediment to a sandy reference
sediment would be conservative, contaminants are going to be bound less tightly to the
sandy sediments, and in a short term study could actually be more available to benthic
organisms.
I am also concerned with the approach of comparing bioaccumulated levels in test
organisms with values reported for organisms in the vicinity of the HARS. In the
example provided in Appendix B, contaminant levels in the comparison data for worms
and clams were quite high, indicative of exposure to contaminants placed previously at
the HARS. Just because bioaccumulation from a candidate sediment does not exceed
these values, it is not necessarily clean enough for placement at the HARS if remediation
is really the goal.
Proposed Additions to Analvte List: Alkylated PAHs
2. Is measurement of the 16 priority pollutant PAHs (i.e. parent PAHs) sufficient for
characterizing the risks associated with the total PAH bioaccumulated by organisms
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exposed to dredged material proposedfor placement at the HARS? Does measurement of
the alkylated compounds significantly improve risk assessment of PAHs?
I do not think measuring only the 16 priority pollutant PAHs is sufficient for
characterizing the risks associated with total PAHs bioaccuraulated by organisms
exposed to dredged material proposed for placement at the HARS. Particularly when
there are significant petroleum sources, nonalkylated (parent) PAHs do not necessarily
represent the major fraction of PAHs in the sediment (reviewed by Irwin et al, 1997).
Given the proximity of refineries, marine terminals, and urban run-off to the NY-NJ
Harbor Complex, it is very likely that sediments from the Harbor proposed for placement
at the HARS would have significant concentrations of alkylated PAHS. Due to their
increased hydrophobicity relative to nonlkylated PAHs, alkylated PAHs would be
expected to be more persistent, to be bioaccumulated to a greater degree, and some
studies on aquatic organisms have shown them to be more toxic. The problem from the
standpoint of human health risk assessment, or even aquatic health risk assessment for
that matter, is the lack of toxicity data at the present time. Therefore, at the moment,
concentrations of alkylated PAHs will need to be treated as being equivalently toxic to
parent PAHs. This may result in either an under or over estimation of their toxicity, but
is certainly preferable to leaving them out all together as is currently done.
3. Is the proposed adaptation of EPA Method 8270 (Appendix D) acceptable and
appropriate for regulatory decision-making? If not, what is an acceptable and
appropriate method?
Although I am not an expert on analytical chemistry, the method proposed (EPA 8270)
seems adequate and appropriate. Not only does it include commonly found aklyated
PAHs, but it also includes some important S-heterocyclic PAHs. The increased detection
limits afforded by this method should solve many of the problems created when non-
detectable values are reported. Use of this method, that was adapted from the methods
developed for use in NOAAs National Status and Trends Program, has the added
advantage that it will generate data comparable to a number of relatively recent nation-
wide surveys. My only concern with this method, is as reported it does not include any
N-heterocyclic PAHs on its analyte list. Speaking from a non-chemist's perspective, I
would hope that this could be relatively easily remedied. Some of the N-heterocyclic
compounds such as 9H-carbzole, 7H-dibenzocarbazole, and acridine are known to be
very mutagenic, and are usually not quantified in environmental samples (Mastrangelo et
al, 1996).
4. Under what specific conditions would the testing for alkylated PAHs for a particular
project be appropriate and warranted?
As discussed above, given the known inputs of hydrocrabons to the Harbor Complex,
alkylated PAHs should be analyzed in all samples. It does not make any sense to me to
allow use of one method for some samples, and another method for others. If PAHs are
going to be quantified in any samples, then the entire suite should be assessed.
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5. What uncertainties would be introduced within the analysis of risk should alkylated
PAHs be included? What steps could be taken to account for these uncertainties in
decision-making? Given the likelihood the method for using non-detects (as described in
EPA/CENAN, 1997) will result in an overestimate of risk, what are the implications?
The major uncertainty involved with including alkylated PAHs in the analyte list is the
lack of information on their relative toxicity, particularly to humans. But, given their
persistence, bioavailability, and the likelihood that they will be present in significant
concentrations, I do not think these limitations should preclude their inclusion in the risk
analysis. Without additional information, I think the only choice available is to treat them
as if they were the corresponding parent PAH for the purposes of risk assessment. It will
not be clear how bad a problem the inclusion of non-detects will be until sufficient data is
available to estimate the occurrence of non-detects in samples from the area. However,
limitated data presented in NOAA's 1996 report on sediment toxicity in the Hudson
Raritan Estuary (NOAA, 1995) indicates measurable levels of at lease some of the
alkylated PAHs in sediments from sites that might be considered for placement at the
HARS. Although the basis on which the method for using non-detects described in
EPA/CENAN, 1997 was not provided, it appears to be conservative in its approach. If
applicants for a permit thought this method was overestimating risk, they could require
their contractors to analyze sufficient material to achieve detectable levels of analytes.
Given the limited information on the relative toxicity of alkylated PAHs, erring on the
side of conservatism in the risk assessment does not seem inappropriate.
Proposed Additions to Analyte List: Orwnotins
6. It is recognized that additional methods have been used for the analysis of organotins
(e.g., Krone et al., 1989). Will the proposed analytical method (Rice et al., 1987) provide
adequate data of sufficient quality to assess relevant risks from organotins? If not, please
provide recommendations.
I do not have the expertise to really comment on this question, nor was the Krone et al.
report provided in our background materials.
7. What special QA/QCprocedures should be implemented to ensure the quality and
usability of the organotin data?
Again, since this is outside my professional experience, and I don't have the information,
I really cannot really comment, other than to caution that my colleagues who specialize in
metals analysis are very skeptical of any data where QA/QC procedures are not reported.
The potential for contamination of reference samples, leading to the conclusion of
insignificant metal enhancement in test samples is something that should be closely
watched when low level enhancement is all that is expected.
8. Under what specific conditions would the testing for organotins for a particular
project be appropriate and warranted?
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Although use of organotin antifouling paints has been banned for use on small (<25 m)
boats since the late-1980s, it is still used on most of the commercial ships using the
Harbor Complex. A recent review on the ecological risks associated with TBT in US
surface waters concludes that water column concentrations have decreased substantially
in most surface waters, although levels particularly in some marinas and a few shipyards
still exceed EPAs chronic water quality criterion of 10 ng/L (Cardwell et al, 1999). There
were no harbors in the NY/NJ complex included in this assessment. However, as
sediments serve as repositories for lipophillic contaminants such as organotins, one
would expect sediment concentrations to remain higher longer, and still serve as a source
of organotins to both the water column and to marine organisms.
Proposed Additions to Analvte List: Coplanar PCB Congeners
9. If the approach for evaluating dioxin is modified, should it include the contribution of
PCBs with dioxin-like activity as proposed? If so, how?
Yes, if dioxin is no longer going to be evaluated using a threshold value, the contribution
of PCBs with dioxin-like activity should be assessed. The most appropriate way to do
this would be to adopt a method that measures the co-planner PCBs that have toxic
equivalency factors assigned to them, and then to include their contribution to dioxin-like
activity when conducting the risk assessment. As co-planner PCBs may be present at
concentrations far exceeding those of the dioxins and furans, their contribution to risk
could be significant.
Comparison to Reference
10. Please consider the policy for assigning values (at one half the detection limit) to
tissue residues that are reported as "
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11. Is the use offunctional groupings in statistical comparisons to reference appropriate
and/or preferable to statistical comparisons using individual contaminants for the
purposes of risk analysis?
Statistics is not my field, so I cannot comment rigorously on this topic. However, unless
it is known that all components of a group act by a similar mode of action, and in an
additive way, summing all members of a functional group is not justified. Having said
that, if there are no appropriate data from which to conduct a risk analysis for all
individual contaminants, one is forced to either remove them entirely from the
calculation, thereby assuming they pose no risk, or add their concentrations to those of
similar compounds for which there is toxicological data. This could either over or
underestimate risk, depending on the compound. In my opinion it would seem better to
include all compounds in the risk assessment even if specific data is not available, rather
than conduct the risk assessment as if they were not there.
Adjustment to Steady State: Organic compounds
I have an issue with this section that was not addressed into the questions posed to the
peer reviewers. That concerns the use of the polychaete N. virens as a model organism to
determine bioaccumulation. We have known for more than 30 years that N. virens
possess the ability to a number of organic contaminants, particularly PAHs. Body
burdens of parent PAHs can be underestimates of total (parent plus metabolite) body
burdens by 90 to 100 fold (McElroy, 1990). Furthermore these metabolites can be
bioaccumulated by consumers (McElroy and Sisson, 1989, McElroy et al, 1991). I do not
feel these risks are being adequately calculated.
12. Is it appropriate to apply a multiplier based on log Kowfor these compounds
(organics), or are there other specific data that can be used to estimate steady state? If
so, please identify.
I agree that this is impractical to require applicants to conduct long-term (>28 day)
bioaccumulation studies to evaluate sediments for placed at the HARS. Using a
multiplier based on log KoW and previous empirical data seems like an appropriate
approach. I do have several concerns with this approach. As discussed in Lee at al
(1994), rates and final steady state values for organic contaminants bioaccumulation are
species specific. I have also reported this in my own work (McElroy and Means 1986;
Means and McElroy, 1997). Organisms that deposit feed aggressively can reach steady
state much more quickly. These differences are not removed by normalizing for lipid
content. There are also species-specific differences in gut residence time and gut
surfactancy that have significant effects on contaminant uptake efficiency (Ahrens et al.
2001, a,b). As there are only a limited number of organisms for which we are concerned
about trophic transfer of contaminants in sediments to be placed at the HARS, it would be
better to have both contaminants class- and species-specific KoW multipliers to estimate
steady state concentrations. The KoW vs. proportion of SS bioaccumulation reached in 28
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days figures (6-1) in EPA/US ACE 1998 does not specify for which organisms this
relationship was determined. The reference cited (McFarland's, 1994 Ph.D. thesis was
not included in the material provided for this review) Finally, as discussed in Lee et al.
(1994), even when studies such as these are conducted identically, large (several-fold)
differences in bioaccumulated values can be obtained. To go with a single set of numbers
largely derived from one study does not seem supportable.
13. Given the increased hydrophobicity of alkylated PAHs, is the use of the correction
factor associated with the corresponding parent an appropriate approach for estimating
steady state residues of alkylated PAHs? If not, please elaborate.
Because of the increased hydrophobicity of alkylated PAHs I do not think that applying
the KoW of the unalkylated PAH to estimate a correction factor for steady state
bioaccumulation of alkylated PAHs is appropriate. With minimal effort scientists at the
EPA or USACE could measure accurate KoWs for these compounds, and conduct long
term bioaccumulation studies conducted on representative members of this group
(including both the heterocylic and alkylated PAHs).
14. For the DDT derivatives and dieldrin, please comment on the appropriateness of
using M. nasuta data rather than N. virens-specific data in the estimation of steady state
multipliers.
Pruell et al. (1990) reported bioaccumulation factors of 0.8 and 0.5 for total PCBs for M
nasuta and N. virens respectively. Based on their results, applying the proposed
multipliers derived for M. nasuta to N. virens would underestimate steady state
concentrations for N. virens. As I mentioned above, I think you need both species and
compounds class specific multiplication factors, and the potential contribution of
metabolites should also be addressed if species that metabolize organic contaminants
such as N. virens are going to be used in bioaccumulation tests. There are other large,
locally abundant polychaetes that have limited ability to metabolize aromatic
hydrocarbons such as the blood worm, Glycera alba (McElroy et al, 2000) that could be
used if data from a polychaete is essential.
15. Are the approaches taken to adjust organic contaminant bioaccumulation data to
steady state adequate? Do the proposed multipliers agree with previously published
studies (i.e., do they appear reasonable)? If not, please elaborate.
Taking into account the considerations listed above (the need for species and compounds
class specific values, and the need to use numbers derived from a number of studies), I
think the KoW approach to adjust organic bioaccumulation data to predict steady state
bioaccumulation is adequate, with one exception. The value reported for p-p'DDT for M.
nasuta of 11 seems way out of line. Lee et al. (1994) did not really discuss this apparent
outlier. Before applying this value, I would want to see further experimental data to
support this number.
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16. What are the major sources of uncertainty associated with the approaches? What
alternative approaches would reduce the uncertainties? How could these uncertainties be
described and accounted for in decision-making?
As already discussed, I think if you have compound-class specific and organism specific
steady state multipliers derived from a number of experimental determinations on field
collected sediments, this approach is acceptable. However, the confounding factor of
using organisms that can metabolize organic contaminants readily, such as N. virens, has
not been adequately addressed.
Adjustment to Steady State: Metals
17. In your opinion, is the methodology followed to derive the steady state multiplier for
non-essential metals (i.e., a factor of three) scientifically appropriate (Appendix G)?
Please elaborate. Do you have any recommendations of additional or alternate
methodologies or information that can be used to either supplement or replace the
proposed method?
I am not an expert on metal bioaccumulation, but I find myself confused by the
discussion in the Proposed TEF concerning steady state bioaccumulation of metals. Why
does a "true" steady state for metals not appear to exist? Battelle's study of metal body
burdens in N. virens collected from in and around the HARS site in 1996 showed
variations of less than three-fold in body burdens with no clear geographic trends. Based
on this information a decision was made to apply a safety factor of 3 to metal data
obtained from 28-day bioaccumulation tests. This seems somewhat arbitrary to me.
Human Health Evaluations: Overall
General Comments: I cannot comment in detail on the following questions as I have no
experience in conducting human health risk assessments. But I do have some general
comments. I find some of the assumptions used in developing these methods to be
somewhat arbitrary, and in reality there is certainly not a single factor appropriate for
estimating trophic transfer of all PAHs, the whole body to filet ratio, seafood
consumption, or the site use factor. It would seem more appropriate to develop a
probabalistic approach that evaluates ranges of these parameters to determine what
contributes most to estimating cancer and non-cancer human health risk.
Trophic transfer factors (using 0.1 for PAHs). This seems low. Part of this problem may
likely be associated with only quantifying unmetabolized parent PAHs when assessing
bioaccumulation. The TEF does not reference where they get their value of 0.1 for
PAHs, except to say that it was obtained from "literature values". In the example risk
assessment given in Appendix B, some of this literature is cited, but after re-reading it
(some of which is my own) I feel that the value of 0.1 is not conservative. My own
experiments examining trophic transfer of BaP from N. virens into winter flounder that
24 hours after an oral does of BaP, somewhere in excess of 20% of the dose can be
recoved in the bile, liver and intestine alone. These experiments were done with
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radiotracers, allowing for the contribution of metabolites body burdens to be included.
Mixed BaP metabolites from worms were transferred approximately 4-fold less
efficiently. (McElroy and Sisson, 1989). Rereading Varanasi's 1989 review on the
subject also indicates wide variability. This problem would be exacerbated by the
presence of metabolites in prey tissue. Levels that are currently not being quantified at
present. For worms such as N. virens, this could inflate total PAH (parent and
metabolites) body burdens by as much as a factor of 100.
18. Please comment on each factor listed above (and in Table 5) as to its appropriateness
for use in the equations listed above. Would you recommend additional factors? Would
you change or modify the equations as written above? If so, how?
I have nothing specific to add to what I've already said above.
19. Are the methods used to derive the human health exposure parameters and assigned
values discussed in Section E appropriate (please review the referenced appendices)? If
not, please elaborate. How should these factors be factored into the risk analyses and
decision-making?
I have no specific comments on this question.
20. Is the approach taken to relate fish whole body and fillet concentrations scientifically
appropriate? If not, what method would you recommend?
Not much of a description of how these factors were derived was provided in the TEF,
particularly for organic contaminants. I cannot speak to the issues with metals, but I
would assume that lipid content would significantly influence how well this relationship
works. Given the relatively limited number of seafood species consumed form the HARS
it might be best to develop species-specific multipliers.
21. Could the analysis be improved by focusing on key fish (seafood) species at the
HARS? What characteristics should be used to select these key species?
As discussed above, I do think the analysis could be improved by focusing on a few key
species at the HARS. They should be species caught in the area, and reflect fish and
shellfish, including both molluscs and Crustacea. Further, species that have regularly
migrate up and down the coast spending time in highly polluted areas such as striped bass
and bluefish should be excluded from the analysis as their migration patterns would
preclude linking contaminant exposure to the HARS. I would recommend, lobster,
surf clam, summer and winter flounder at a minimum. All are resident in the HARS
area year-round, have limited migratory excursions, and are harvested in both the
commercial and recreational fishery.
22. In your opinion, is the approach for assuming total metal to be in the most toxic form
appropriate and reasonable? Should metal speciation/complexation be considered in the
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assessment of metals bioaccumulation, trophic transfer, and human health risks? Is the
proposed approach for evaluating methyl mercury appropriate? Are there alternative
analytical or risk assessment techniques available that would improve the risk
assessment of metals? Is the multiplier proposedfor adjusting measured concentrations
of arsenic appropriate and reasonable?
I do not have the background to address this question.
23. Is the assumption that the potency of alkylated PAHs can be estimated by the potency
of the parent PAH appropriate? Is this assumption likely to result in an under- or
overestimate of the risk associated with the alkylated PAHs?
As there are no human health data on the potency of alkylated PAHs this is a difficult
question. Using the potency of the parent PAH is better than not including them at all in
the risk assessment. However, there is no way to know without data whether or not this
is under or overestimating their risk.
24. Please comment on the potential for human exposure to PAHs through consumption
of finfish and other seafood.
There is the potential for human exposure to PAHs through the consumption of finfish
and other seafood. Due to their extensive ability to metabolize and eliminate PAHs, the
potential exposure from consuming finfish is probably small. The risks associated with
consuming crustaceans and mollusks is much larger due to their limited ability to
eliminate these compounds.
25. What are the major sources of uncertainty associated with the approaches described
in Section E? What alternative approaches would reduce the uncertainties? How could
these uncertainties be described and accounted for in decision-making?
As discussed above, using fixed numbers for all the parameters in the risk assessment is
problematic. I would prefer to see ranges provided for most of these, and some sort of
probabalistic modeling used that take this into account.
26. What is your recommendation for evaluating the potential toxicity of organotins?
Should they be evaluated as individual compounds? Summed as total? Should there be
some consideration of relative toxicity?
I am not familiar enough with the relative potency of the individual organotin compounds
to comment on this question.
27. Please comment on the appropriateness of the proposed approach for converting and
using the analytical data for alkylated and parent PAHs to estimate risk from all PAHs.
This question has already been addressed.
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Human Health Evaluations: Comparison to HARS-Specific Values
28. Do you believe that the "disaggregate " modeling discussed above (and shown in
Figure 4) for estimating human health HARS-Specific Values for lead is appropriate?
Would you recommend an alternative risk assessment method be used given the
information and data available? Do you believe the method described has appropriately
taken uncertainty into account? Please elaborate.
The method described does not appear to take uncertainty into account at all.
Human Health Evaluations: Consideration of Combined effects
29. In your opinion, are the methodologies and equations described above appropriate
for estimating total carcinogenicity and combined non-cancer impacts of contaminant
mixtures accumulated from dredged materials proposed for use as Remediation Material
at the HARS?
The approach as described in the TEF would seem to be adequate, but inclusion of a
probabalistic approach that would include uncertainty in the variable would provide more
useful information, and a better understanding of the risks involved.
30. Is the conceptual model for evaluating fish exposure to dredged material at the HARS
and human exposure through ingestion of seafood appropriate and reasonable? How can
the uncertainties associated with the assumptions in this conceptual model be reduced?
Please consider the spatial and temporal elements of exposure in your discussion.
Risk assessment is not my field of specialty. I think conceptually, evaluating risk of
human health exposures from seafood caught in or around the HARS is appropriate.
However, given the large uncertainties in the all the parameters used to calculate this risk,
taking a more probabilistic approach to the modeling would seem more appropriate.
Exposure will vary among individual people, among individual species, and over time. It
seems to me that varying these parameters in the risk assessment is the only way to obtain
an estimate as to how much each affects the outcome. I do think focusing on risk
associated with consuming specific species likely to be accumulating significant portions
of their contaminant body burdens from contaminants associated with HARS sediments
is appropriate. I am also concerned about the 10"4 cancer risk level being used in this
assessment. As described in the sample risk assessment provided in Appendix B,
acceptable risk varies between a value of 10"4 and 10"6. The discussion implies that
since they are trying to use a conservative approach in the estimation of risk that
accepting a higher level of risk is still protective. Given the uncertainty in all the factors
used to calculate risk, is this prudent?
References: In addition to those already provided:
Ahrens, M.J. Hertz, J., Lamoureux, E.M. Lopez, G.R., McElroy, A.E. and Brownawell,
BJ.
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2001.The effect of body size on digestive chemistry and absorption efficiencies of food and
sediment-bound organic contaminants in Nereis succinea (Polychaeta). J. Exp. Mar. Bio.
Ecol. 263:185-209.
Ahrens, M.J. Hertz, J., Lamoureux, E.M. Lopez, G.R., McElroy, A.E. and Brownawell,B.J.
2001.The role of digestive surfactants in determining bioavailability of sediment-bound
hydrophobic organic contaminants to two deposit-feeding polychaetes. Mar. Environ.
Prog. Ser. 212:145-157.
Cardwell, R,D, Brancato, MS., Toll, J., DeForest, D., Tear, L. 1999. Aquatic Ecological
risks posed by tributyltin in University States surface water : pre-1989 to 1996 data. Env.
Toxicol, and Chem. 14:567-577.
Irwin, R.J., VanMouwerik, M., Stevens, L., Seese, M.D., and Basham, W. 1997.
Enviornmental COntminants Encyclopedia. National park Service, Water Rsources
Division, Fort Collins, CO. a
Mastrangelo, G., Fadda, E., and Marzia, V. 1996. Polycyclic aromatic hydrocarbons and
cancer in man. Environ. Health Perspect. 104:1166-1170,
Means, J.C. and A.E. McElroy, 1997. Bioaccumulation of PCBs: sediment and animal
variables. Environ. Toxicol. Chem. 16:1277-1286.
McElroy, A., Leitch, K., and Fay, A. 2000. A survey of in vivo benzo[a]pyrene
metabolismin small benthic invertebrates. Mar. Environ. Res. 50:33-38.
McElroy, A.E., Cahill, J.M., Sisson, J.D., and K.M. Kleinow. 1991. Relative bioavailability
and DNA adduct formation of benzo[a]pyrene and metabolites in the diet of the winter
flounder. Comp. Biochem. Physiol. 100C:29-32.
McElroy, A.E. 1990. Polycyclic aromatic hydrocarbon metabolism in the polychaete
Nereis virens. Aquat. Toxicol. 18:35-50.
McElroy, A.E. and Sisson, J.D. 1989. Trophic transfer of benzo(a)pyrene metabolites
between benthic organisms. Mar. Environ. Res. 28:265-269.
McElroy, A.E., and Means, J.C. 1988. Factors affecting the bioavailability of
hexachlorobiophenyls to benthic organisms. Aquatic Toxicology and Hazard Assessment:
10th Volume, ASTM SPT 971, W.J. Adams, G.A. Chapman, and W.G. Landis, Eds.,
American Society for Testing and Materials, Philadelphia, pp. 149-158.
NOAA, 1995. Magnitude and Extent of Sediment Toxicity in the Hudson-Raritan
Estuary. NOAA Technical memorandum NOS ORCA 88. Silver Spring MD.
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Comments Received from
James Meador, Ph.D.
March 21,2002
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SPECIFIC CHARGES FOR SCIENTIFIC PEER REVIEWERS
Comments from James Meador, NOAA Fisheries
Overall Process
1. Throughout the proposed process, there are various uncertainties
introduced. Please identify the key areas of uncertainty that need to
be addressed. Are there additional data sources or parameters that
could be used to address these areas? What methods are available for
describing and accounting for these uncertainties in the calculation of
HARS-Specific Values? Of the methods available, which would you
recommend for consideration and why? Please consider the
implications of implementing these methods in the regulatory
framework. Please include an evaluation of probabilistic and
deterministic methods in your discussion.
For me, the key areas of uncertainty are 1. Consideration of
contaminants that lack solid toxicological data that can be used in the risk
assessment (e.g., alkylated PAHs). 2. Selection of species and
characterization of their site usage. 3. Comparison to reference values and
how to deal with low concentrations. 4. Estimation of steady state for many
compounds. Some of these are relatively minor (e.g., number 4) and may
be dealt with easily. Others, such as number 1 will be much more difficult.
Many assumptions are made here, some that are very conservative and
others that are backed by insufficient data needed to gauge their
appropriateness. I believe it is appropriate to be conservative when there is
enough information to support such a decision, however, for those cases
where the data are incomplete and the uncertainty is too large, a more
detailed analyses is required.
Most of my specific recommendations are listed below in the responses
to individual charges. As stated below, I would like to see some
incorporation of probabilistic methods in this process because boundaries on
individual parameters can be determined and uncertainty can be addressed.
I would also like to see more use of upper or lower percentiles for some
parameters (where appropriate), instead of means or medians to ensure
protection of sensitive individuals. We use this approach under the U.S.
Endangered Species Act and it would also be prudent for human populations.
In general, I am not familiar with many of the methods used in risk
assessment, so I will defer to the panel's recommendations on some of these
issues.
1
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Proposed Additions to Analvte List: Alkylated PAHs
2. Is measurement of the 16 priority pollutant PAHs (i.e. parent
PAHs) sufficient for characterizing the risks associated with the total
PAH bioaccumuiated by organisms exposed to dredged material
proposed for placement at the HARS? Does measurement of the
alkylated compounds significantly improve risk assessment of PAHs?
I would say that measuring just the 16 parent PAHs is not sufficient for
characterizing risk to humans based on total PAH bioaccumuiated by
organisms exposed to dredged material. The addition of several alkylated
PAHs would add to the EPAs ability to adequately characterize risk from PAH
exposure. Several studies have shown that some of these compounds are
mutagenic. For example, LaVoie et al. (1983) tested a series of alkylated
phenanthrenes and assayed for mutagenic activity in Salmonella
typhimurium. The alkylated phenanthrenes assayed, 1-
methylphenanthrene, 9-methylphenanthrene, 1,4-dimethylphenanthrene
and 4,10-dimethylphenanthrene were active as mutagens. Additionally,
some alkylated PAHs, such as methylanthracene are known to become more
toxic when exposed to UV wavelengths (Boese et al. 1998). This is one of
those cases where some information is available to suggest that
toxicologically important compounds are not being considered; however, the
data are not sufficient for including them in any quantitative risk assessment
for human consumption. I would be willing to support a procedure for
incorporating these data if a logical, mechanistic approach is presented.
3. Is the proposed adaptation of EPA Method 8270 (Appendix D)
acceptable and appropriate for regulatory decision-making? If not,
what is an acceptable and appropriate method?
I had our chemists here at NOAA examine the methodology in
Appendix D and they concluded that these methods are appropriate. The
chemists in our group have been analyzing alkylated PAHs by GC/MS for
about 10 years and were responsible for developing some of these
techniques. There are just a few comments. Why are decalins listed in
Table 1? These are not PAHs. Bullet number 3 in section 1.2 (page 88)
"biological tissues exposed to contaminated sediment generally have upper
bioaccumulation potentials of approximately 100 - 500 ppb" is not true.
Table 3 the DQO targets should be 60% - 120% recovery.
2
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4. Under what specific conditions would the testing for alkylated PAHs
for a particular project be appropriate and warranted?
Under any circumstances where petroleum or its products are
expected to be found. Determination of just the 16 parent PAH compounds
implies that only combustion type sources (e.g., soot, automobile exhaust,
or coal). However, these parent compounds are often minor components in
petroleum and related products. Any industries on waterways around the
New York/New Jersey area that use petroleum or petroleum products may
be sources of alkylated PAHs to the environment. Alkylated PAHs are
commonly found in most urban areas (e.g. see the NOAA Mussel Watch
data). It would be appropriate under most circumstances to determine the
extent of contamination by alkylated PAHs; however, as stated in charge 2
above, the procedures for incorporating these into the risk assessment for
human health are likely not tenable.
5. What uncertainties would be introduced within the analysis of risk
should alkylated PAHs be included? What steps could be taken to
account for these uncertainties in decision-making? Given the
likelihood the method for using non-detects (as described in
EPA/CENAN, 1997) will result in an overestimate of risk, what are the
implications?
Given that some alkylated PAHs are mutagenic, I am not sure how
their potency would be determined and applied to a risk assessment.
Something akin to the Toxicity Equivalency Factors (TEFs) that were
generated for PCBs or cancer slope factors would be needed. There appears
to be few if any alkylated PAHs in the IRIS database. Before alkylated PAHs
are added, it would be prudent to search the literature to see if generating
cancer potency factors may be possible.
The approach proposed for HARS specific values that uses EPA (1993)
and substitutes the parent compound value for the alkylated homologue
requires additional information to reduce uncertainty. If the alkylated
homologues are metabolized to the parent compounds for which data are
available, then this may a reasonable approach. Supporting information
about the disposition of the alkylated compounds in fish and humans would
be needed to support this approach and reduce uncertainty.
Proposed Additions to Analvte List: Oraanotins
6. It is recognized that additional methods have been used for the
analysis of organotins (e.g., Krone et al., 1989). Will the proposed
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analytical method (Rice et a I., 1987) provide adequate data of
sufficient quality to assess relevant risks from organotins? If not,
please provide recommendations.
I have compared the Krone et al. (1989) method and the Rice et al.
(1987) method and although they are similar, some important differences
are evident. First, the Rice method was developed for tributyltin, not
organotins. This method does not use tropolone, which is important for
extracting dibutyltin and monobutyltin. The Krone method uses tropolone.
Without tropolone, extraction efficiencies for these other butyltins would be
very poor. The addition of copper, which is not used in the Rice method but
is used in the Krone method, is important in diminishing sulfur compounds
that interfere with the flame photometric signal. The addition of copper to
sediment improves the detection limits considerably. The Krone method is
used by analytical labs. In our area I know of two commercial laboratories
that use the Krone method for organotin analysis. One is a very large
laboratory with offices around the country (Columbia Analytical Services).
7. What special QA/QC procedures should be implemented to ensure
the quality and usability of the organotin data?
For organotin analysis, appropriate recovery standards, GC internal
standards and certified reference material (CRM) are required. Additionally
method blanks, spiked blanks, and calibration standards also contribute to
high quality analyses. For example, in the method of Krone et al. (1989),
tripentylmonobutyltin was employed as GC internal standard. In this
method, calibration was done by the internal standard method using peak
heights and a calibration curve with five concentrations. For the recovery
standard, tripentyltin chloride is commonly used. Additionally, TBT
determinations should be adjusted for the recovery of tripentyltin, which is
run with each set. The limit of detection for tissue and sediment should be
close to 5 ng/g (dry weight).
For many years a tissue CRM (seabass) has been available to gauge
accuracy of tributyltin and triphenyltin (Okamoto, 1991). The certified value
is certified at 1.3 (0.1) pg/g (mean and standard deviation). The seabass
CRM is called NIES CRM number 11 and it can be obtained from the location
listed below in the citation section. More recently, a sediment CRM has
become available from the National Research Council of Canada, who
supplies a large number of CRMs for various chemicals and matrices. The
sediment CRM is called PACs II.
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8. Under what specific conditions would the testing for organotins for a
particular project be appropriate and warranted?
I would say under all conditions where sediment is being dredged from
harbors. Most harbors and ports around the world are contaminated with
organotins from shipping and drydock activities. The Organotin Antifouling
Paint Control Act of 1988 (OAPAC; US Congress 1988) banned TBT paint on
vessels less than 25 m. A large percentage of ships longer than 25 m are
still using TBT antifouling paint because it is so highly effective at controlling
biofouling on ship hulls. Because organotins are fairly persistent in
sediment, areas where recreational craft predominate generally contain
elevated concentrations of organotins in sediment. I would also consider
analyzing for phenyltins. These are also immunotoxic organotins that occur
often in high concentrations in urban areas. Triphenyltin has been used in
some antifouling paints and as fungicides in agricultural areas.
Proposed Additions to Anaivte List: Coplanar PCB Congeners
9. If the approach for evaluating dioxin is modified, should it include
the contribution of PCBs with dioxin-like activity as proposed? If so,
how?
In recent work it has been shown that some PCB congeners are much
more toxic than others, which is primarily a function of the position of the
chlorine atoms and their ability to interact with the aryl hydrocarbon (Ah)
receptor. This is mainly a concern for vertebrates, including fish, because
invertebrates generally lack this receptor and are not usually sensitive to
the dioxin-like effects of PCBs.
The most toxic PCBs are the non-ortho and mono-ortho substituted
congeners, which tend to be planar compounds. Some of these responses
listed above, such as developmental and reproductive abnormalities,
enzyme induction, and immunosuppression, can occur at extremely low
concentrations and are likely caused by "dioxin-like" PCB congeners (planar
congeners).
The Toxicity Equivalent Factor (TEF) approach has been used to
determine the relative toxicity of the planar PCB congeners as a fraction of
that elicited by 2,3,7,8 tetrachlorodibenzo-p-dioxin (TCDD). Tissue
concentrations of PCB congeners are multiplied by the TEF to generate a
Toxicity Equivalent (TEQ) concentration in terms of its "dioxin-like" potency.
These TEQs are then summed to generate a total TEQ concentration for the
sample that can be compared to dioxin toxicity results. TEFs have been
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developed for mammals and birds and only very recently for fish (Van den
Berg et al. 1998). PCBs that elicit dioxin like effects should be included,
which is a relatively simple matter. Even though such toxic congener may
occur at low concentrations, they must be considered for an accurate
human health risk assessment.
Additionally, the TEF approach is not applicable for those "non-dioxin-
like" biological responses caused by the nonplanar PCB congeners, primarily
due to the different modes of action. The responses caused by the non-
planar congeners ("non-dioxin-like") are likely due to different modes of
action and include neurotoxicity, hypothyroidism, carcinogenicity, behavioral
alteration, and endocrine disruption (Giesy and Kannan 1998). These non-
dioxin like modes of action for PCBs should be included in any risk
assessment.
Comparison to Reference
10. Please consider the policy for assigning values (at one half the
detection limit) to tissue residues that are reported as "
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estimating the mean and variance of the data (Gilbert 1987, Newman et al.
1989). A comparison of mean and variance values with and without the
correction showed substantial differences, especially when the number of
below detection values was high.
A PC program called "UNCENSOR" has been written to deal with left-
censored data and it is available from Dr. Michael Newman at the Virginia
Institute of Marine Sciences (VIMS).
11. Is the use of functional groupings in statistical comparisons to
reference appropriate and/or preferable to statistical comparisons
using individual contaminants for the purposes of risk analysis?
It would seem appropriate for total PCBs and total DDTs because the
reference dose and cancer potency factors are stated as such for these
groups (Table 5 in the binder); however, I believe the TEQ approach for PCB
congeners and dioxins should be included.
I see this as two different steps; one being the bioaccumulation
comparison and the other a determination of risk. Even if the individual
contaminant measured in tissue from the site is not statistically higher than
that measured for the reference, why not run the risk analysis? If the tissue
concentration approaches a level of concern, then there could be a problem.
It should have nothing to do with comparison to a reference site, which may
also contain contaminants at elevated levels. As for comparing individual
compounds, I think an additivity model, such as TEQs, should be considered
for PAHs. If all PAH RfDs are based on the same toxicity response, then
they could be added up in a similar fashion as that for dioxins and PCBs.
Adjustment to Steady State: Organic compounds
12. Is it appropriate to apply a multiplier based on log Kow for these
compounds (organics), or are there other specific data that can be
used to estimate steady state? If so, please identify.
The rate of elimination (k2) of a compound determines when steady-
state tissue concentrations will occur. Based on the scant literature
available, it is clear that the k2 is often quite variable, even within a study
for a given species. To me, the Kow is just a surrogate for determining
steady state, which in some cases may be useful. In Boese and Lee (1992)
a table listing several species and compounds was generated showing k2
values and predicted values based on Kow. In many cases, there was a
large divergence between measured and predicted k2 values. The best
information for determining steady state tissue concentration would come
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from measured k2 values generated from toxicokinetic experiments
performed under varying conditions. For Macoma nasuta and Nereis virens,
some k2 values for various organic compounds have been determined (see
Boese and Lee 1992 for a review). I would prefer to see an analysis of all
existing data for these species before the multiplier approach is taken.
These types of toxicokinetic experiments are not difficult to perform. Any
lab that conducts the bioaccumulation or toxicity studies could easily
generate this information especially as a subset of required bioaccumulation
testing.
For those studies where some detailed information is available, all
values must be used to determine the lowest value (e.g., 10th percentile)
that would appropriate for a conservative prediction of the time to steady
state. As seen in Boese and Lee (1992) there is considerable variability,
even for species that are suspected of not having any substantial ability to
metabolize these compounds.
13. Given the increased hydrophobicity of alkylated PAHs, is the use
of the correction factor associated with the corresponding parent an
appropriate approach for estimating steady state residues of alkylated
PAHs? If not, please elaborate.
As stated elsewhere in this response, the Kow correction factor
approach would be appropriate only when measured values of k2 are not
available. Based on the data provided in Boese and Lee (1992) showing
measured and predicted k2 values for Macoma nasuta over a range of PCB
congeners, it appears the Kow correction may be a reasonable
approximation. I realize that there are very few k2 values for alkylated
PAHs in the literature, so approximating steady state with the Kow approach
will allow this analysis to continue until such values are generated.
It was stated on page 16 in the proposed TEF that the proposal is "to
adjust residues of alkylated PAHs by the same multiplier as its parent
(unalkylated) homolog". This would be unacceptable. For each compound,
the Kow should be found and the multiplier determined as described for the
other compounds in Table 3 of the binder. Some of these Kows are available
in Mckay et al. (1992) and others in the general literature. For those where
no Kow can be found, some approximation will be needed. With the
availability of models (e.g., SPARC) a Kow can be generated for almost any
organic compound.
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14. For the DDT derivatives and dieldriri, please comment on the
appropriateness of using M. nasuta data rather than N. virens-specific
data in the estimation of steady state multipliers.
In general, I would expect M. nasuta to posses less metabolic
capabilities for organic contaminants compared to Nereis virens. However,
exceptions due occur and one was found in Pruell et al. 1990, who showed
no elimination of PCBs in Nereis virens. As stated in Pruell et al. (1990),
there have been previous studies with polychaetes that have shown variable
results. As cited by Pruell et al. (1990), one study (Fowler et al. 1978)
reported "significant" depuration by Nereis diversicolor, but other studies
(McLeese et al. 1980 and Lake et al. 1985) found no elimination of PCBs
held for 21 to 28 days for Nereis virens. There is specific information
available for Nereis virens, which indicates k2 elimination rates ranging from
0.024 to 0.26 (d-1) for DDT and 0.01 to 0.024 (d-1) for dieldrin (Haya and
Burridge 1988, also see Boese and Lee (1992). Again, where possible,
species-specific information would be best.
15. Are the approaches taken to adjust organic contaminant
bioaccumulation data to steady state adequate? Do the proposed
multipliers agree with previously published studies (i.e., do they
appear reasonable)? If not, please elaborate.
Adjusted to steady state in this general way may work for compounds
that are not metabolized and are fairly hydrophobic (however see answer to
Charge 12 above). The problem is that elimination of a contaminant is really
species specific, (as noted above it is the rate of elimination that determines
steady state.) For many compounds, the rate of elimination varies widely
among species and application of a general multiplier is not appropriate.
Some compounds that are metabolized may come to steady state very
quickly. In those cases, application of the multiplier would be overly
conservative. In other cases, for example tributyltin, steady state may take
75 days for some species, even though it is being metabolized. Application
of a multiplier based on the Kow for TBT would fail miserably in predicting
steady state. The multiplier approach using Kow should be used only in the
absence of real data that include rates of elimination.
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16. What are the major sources of uncertainty associated with the
approaches? What alternative approaches would reduce the
uncertainties? How could these uncertainties be described and
accounted for in decision-making?
The empirical approach using measured k2 values, which has a certain
degree of uncertainty that needs to be included. As seen in Table 6.3 in
Boese and Lee, substantial variability exists for measured k2 values for a
given species and contaminant, even within a study. For example k2 values
for Nereis virens and DDT varied 10 fold in the study by Haya and Burridge
(1988). In cases such as these, all data should be examined for quality
control problems and those data that pass scrutiny, would then be
summarized with a mean and standard error. Choosing a low percentile
value may be prudent to cover all situations.
The following is an example of why previous data should be checked.
Table 4 in the HARS proposed TEF lists the fraction of steady state for
Macoma nasuta and total PCBs at 81% after 28 days and this is based on
data from Pruell et al. (1990). In Boese and Lee (1992) the total PCBs k2
listed for Macoma nasuta is 0.012, which means after 28 days the tissue
concentration should be only 28.5% of steady state. This value was cited as
coming from Pruell et al. 1990. Examination of Figure 6 in Pruell et al.
(1990) shows rapid elimination of total PCBs and I calculated a k2 of 0.112.
This k2, would produce 95% of steady state residues after 28 days. The
uptake curve, is however different showing a large range in values at day
28. This curve indicates steady state from 50% to approximately 95% after
28 days. The main point here is that variability exists in the literature due to
many factors, including calculation errors and individual variability.
Multipliers such as these should be carefully selected and supported with
more information detailing the observed variability.
Adjustment to Steady State: Metals
17. In your opinion, is the methodology followed to derive the steady
state multiplier for non-essential metals (i.e., a factor of three)
scientifically appropriate (Appendix G)? Please elaborate. Do you have
any recommendations of additional or alternate methodologies or
information that can be used to either supplement or replace the
proposed method?
Frankly, I don't understand how this analysis will lead to factor for
assessing steady state. This particular analysis has no temporal component,
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so any connection between tissue residues observed at 28 days and steady
state seems tenuous. I would take this data set in Appendix G and generate
BAF values for each element and species that would approximate those at
steady state. (Because they are field values and supposedly the organisms
were exposed for long periods of time.) I would then generate a dataset of
28 day BAFs from the many bioaccumulation tests run for permit application
or from the literature and conduct a statistical comparison. The steady-state
multiplier could then be determined by simply calculating the ratio of the
median values. Determining the variance of each dataset would be useful
for gauging uncertainty.
Human Health Evaluations: Overall
18. Please comment on each factor listed above (and in Table 5) as
to its appropriateness for use in the equations listed above. Would
you recommend additional factors? Would you change or modify the
equations as written above? If so, how?
Why is arsenic the only element listed with a cancer potency factors?
Several other metals (chromium, nickel, cadmium, and possibly lead) are
known human carcinogens. Shouldn't they be included?
19. Are the methods used to derive the human health exposure
parameters and assigned values discussed in Section E appropriate
(please review the referenced appendices)? If not, please elaborate.
How should these factors be factored into the risk analyses and
decision-making ?
I don't have expertise in this area, so no comments are offered. I will
defer to the panel's recommendation.
20. Is the approach taken to relate fish whole body and fillet
concentrations scientifically appropriate? If not, what method would
you recommend?
It wasn't clear as to why this factor was needed. More detail should be
provided regarding its use.
There are no data presented in Appendix K for organics making any
assessment of scientific appropriateness difficult. For most of the organics,
the controlling factor will be lipid content. For some species, the lipid
content in muscle can be highly variable over life stage and seasons. For
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example, salmonids exhibit highly variable lipid content in muscle, which
may also be true for other species that contain high fat content.
As for metals, the data are incomplete. One paper really isn't sufficient
for making generalizations. Based on the data presented for metals, a
conservation assumption of 1:1 may be reasonable.
21. Could the analysis be improved by focusing on key fish (seafood)
species at the HARS? What characteristics should be used to select
these key species?
Fish with high lipid content should be selected. Also, shellfish such as
lobsters and clams may contain higher concentrations of most contaminants
because of their close association with the benthos and limited metabolic
capacity.
22. In your opinion, is the approach for assuming total metal to be in
the most toxic form appropriate and reasonable? Should metal
speciation/complexation be considered in the assessment of metals
bioaccumuiation, trophic transfer, and human health risks? Is the
proposed approach for evaluating methyl mercury appropriate? Are
there alternative analytical or risk assessment techniques available
that would improve the risk assessment of metals? Is the multiplier
proposed for adjusting measured concentrations of arsenic appropriate
and reasonable?
Assuming that the total metal is the most toxic form would be a
conservative approach. Most elements occur in several forms that range
widely in toxicity. For example, inorganic arsenic is believed to be much
more toxic than the organo forms. In fish, inorganic arsenic is generally
about 10% of the total arsenic, although some species contain higher
percentages. Another example is mercury and selenium, which form an
insoluble complex and likely renders the mercury nontoxic. One important
factor is the fate of such compounds that are ingested and if those elements
can be converted to their toxic forms. Unless this question can be answered
for each element, assuming the total metal is the most toxic form may be
reasonable, but likely conservative.
If possible, it would also seem reasonable to include inputs from other
sources for some metals such as mercury and arsenic, e.g., as was done for
lead. If current data exist on the concentrations in the general population
for these elements, then the incremental amount ingested from fish caught
at the HARS could be considered in the overall risk assessment.
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Is the proposed approach for evaluating methyl mercury appropriate?
This question was not in the original question 22 found in the binder and I
didn't see any mention of an approach for evaluating methyl mercury.
Where was this presented?
23. Is the assumption that the potency of alkylated PAHs can be
estimated by the potency of the parent PAH appropriate? Is this
assumption likely to result in an under- or overestimate of the risk
associated with the alkylated PAHs?
As stated in charge 5, some of the alkylated PAHs are mutagenic and
should be considered; however I am not sure about the best approach for
utilizing this information. For some of the alkylated homologues, using the
parent compound may be appropriate; however, I will defer to the panel's
recommendation.
24. Please comment on the potential for human exposure to PAHs
through consumption of fin fish and other seafood.
It is well known that fish generally have active metabolic systems and
can metabolize PAHs as they are accumulated. This results in generally low
tissue concentrations of these compounds. However, examination of the bile
offish exposed to PAHs shows very high concentrations of PAH metabolites,
which have been changed from the parent compound to render them more
water soluble. For those people who would ingest whole fish, uptake of
these metabolites may be an important source of mutagenic compounds.
Shellfish, such as clams and lobsters are known to have very weak or
non-existent abilities to metabolize PAHs and are able to accumulate these
compounds to very high levels. Ingestion of shellfish should be an important
consideration of human exposure to PAH compounds. Bioaccumulation
factors and trophic transfer factors should be generated for shellfish and
considered in human risk assessment.
25. What are the major sources of uncertainty associated with the
approaches described in Section E? What alternative approaches would
reduce the uncertainties? How could these uncertainties be described
and accounted for in decision-making?
For most models, a sensitivity analysis is usually performed to
determine which factors have the biggest influences. Performing such an
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analysis will help define the range for each value and the overall effect it will
have on the result.
26. What is your recommendation for evaluating the potential
toxicity of organotins? Should they be evaluated as individual
compounds? Summed as total? Should there be some consideration
of relative toxicity?
Human risk assessment with butyltins is mostly concerned with
tributyltin, which has been shown to be toxic to the thymus-dependent
immune system. The EPA has set a reference dose for TBT at 0.3
ug/kg/day; however, there are other published values ranging from 0.25 to
6.7 ug/kg/day (see Cardwell et al. for a review). Dibutyltin has also been
shown to be thymotoxic in laboratory studies (Seinen 1980) and this
compound is also known to inhibit oxidative phosphorylation and is generally
cytotoxic (see review by Snoeij et al. 1987). I didn't find any mention of
monobutyltin in these studies of organotin toxicity, so it is not known if this
compound will produce similar toxicity in humans. Tetrabutyltin, if it occurs,
should also be included because it will be metabolized to tri- and dibutyltin.
Note: Table 5 lists the RfD as 0.3 mg/kg/day for TBT, which is 1000 times
higher than the accepted value.
I would recommend treating tetrabutyltin, tributyltin, and dibutyltin
together in a human health risk assessment. They could be considered
equally toxic, which would allow use of the RfD set for TBT, or it may be
possible to derive the RfD for dibutyltin from the literature and apply it to
the observed data.
27. Please comment on the appropriateness of the proposed
approach for converting and using the analytical data for alkylated and
parent PAHs to estimate risk from all PAHs.
Please see my response to charge 23. I will defer to the panel's
recommendation on this charge.
Human Health Evaluations: Comparison to HARS-Specific Values
28. Do you believe that the "disaggregate" modeling discussed
above (and shown in Figure 4) for estimating human health HARS-
Specific Values for lead is appropriate? Would you recommend an
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alternative risk assessment method be used given the information and
data available? Do you believe the method described has
appropriately taken uncertainty into account? Please elaborate.
I thought the approach was logical and reasonable; however, some
conservative assumptions were made. For example, the trophic transfer of
lead from invertebrates to fish may be rather high, based on the data in
Appendix L. As for dealing with uncertainty, it would be useful to examine
data on blood levels in people from the New York/New Jersey area (I would
assume such data exist) and tissue concentrations found in fish from the
HARS.
Human Health Evaluations: Consideration of Combined effects
29. In your opinion, are the methodologies and equations described
above appropriate for estimating total carcinogenicity and combined
non-cancer impacts of contaminant mixtures accumulated from
dredged materials proposed for use as Remediation Material at the
HARS?
I don't have any experience in this area and will defer to the panel's
recommendation. I would like to know how some compounds will be
handled in the risk assessment. For example, some PAHs have Water
Quality Criteria values for human consumption of water and organisms. This
implies that at a water concentration above the criterion, organisms should
not be ingested. To determine this water value, some assessment of
bioaccumulation and human ingestion must have been conducted. In the
EPAs list of WQC, several PAHs have Human Health criteria that are not
listed in Table 5 of the binder. Many of the criteria are high water
concentrations; however, some are very low (e.g. chrysene at 0.004 ug/L).
If there are human toxicity data for these compounds, shouldn't they be
included in Table 5?
30. Is the conceptual model for evaluating fish exposure to dredged
material at the HARS and human exposure through ingestion of
seafood appropriate and reasonable? How can the uncertainties
associated with the assumptions in this conceptual model be reduced?
Please consider the spatial and temporal elements of exposure in your
discussion.
A sensitivity analysis of the proposed model would help identify the
degree of uncertainty in each factor. Running the model with the factors
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that are the most variable and examining the model outcome for extreme
values will help identify those factors with the biggest influence.
As for the temporal elements, it appears that the steady-state
adjustment factors are generally reasonable. I assume the spatial elements
refer to habitat usage by fish, which to me is probably one of the most
variable components of this model. The data presented in Appendix J was
used to generate an average site use for all species of fish caught at the
HARS, which is based on commercial catch data. One variable is the fishing
effort throughout the year. Are commercial fisheries equally active during all
seasons? Another important factor is species usage. It would be expected
that some species spend much more time at the HARS than others. Because
the surrounding area is sand, the HARS may be rich with food items for
benthic or epibenthic fish. Conversely, fish that are more pelagic (e.g.,
bluefish) probably range widely over the New York Bight, unless they are
attracted to the HARS by available prey.
If a modeling effort is to be used, I would prefer to see a probabilistic
approach such as the one outlined in von Stackelberg et al. (Supplemental
material). To me, an approach such as this is highly defensible because
each parameter has a distribution making it highly amenable to sensitivity
analysis. Unless the uncertainties are quantified in any modeling effort, the
accuracy will always be in question.
I would prefer to see more direct measurements of contaminant
concentrations in fish from the HARS, which could be used to calibrate the
chosen model and provide more accurate information about actual exposure.
Citations
Boese BL, J.O. Lamberson, RC Swartz, R. Ozretich, and F. Cole. (1998).
Photoinduced toxicity of PAHs and alkylated PAHs to a marine infaunal
amphipod (Rhepoxynius abronius). Arch Environ. Contam. Toxicol.
34: 235-240.
Boese, B.L. and H. Lee II. (1992). Synthesis of methods to predict
bioaccumulation of sediment pollutants. U.S. EPA. ERL-N contribution
N232.
Cardwell, R.D., J.C. Keithly, and J. Simmonds. (1999). Tributyltin in U.S.
market-bought seafood and assessment of human health risks.
Human and Ecological Risk Assessment 5:317-335.
Cohen AC Jr (1959). Simplified estimators for the normal distribution when
samples are single censored or truncated. Technometrics 1:217-237
Cohen AC Jr (1961). Tables for maximum likelihood estimates: singly
truncated and singly censored samples. Technometrics 3:535-541
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Giesy, J.P. and K. Kannan. (1998). Dioxin-like and non-dioxin-like toxic
effects of polychlorinated biphenyls (PCBs): implications for risk
assessment. Crit. Rev. Toxicol. 28: 511-569.
Gilbert RO (1987). Statistical methods for environmental pollution
monitoring. Van Nostrand Reinhold Co. New York, NY 320 pp
Haya, K. and L.E. Burridge. (1988). Uptake and excretion of organochlorine
pesticides by Nereis virens under normoxic and hypoxic conditions.
Bull. Environ. Contam. Toxicol. 40: 170-177.
LaVoie EJ, Tulley-Freiler L, Bedenko V, Hoffmann D. (1983). Mutagenicity
of substituted phenanthrenes in Salmonella typhimurium. Mutat Res
116(2):91-102
Newman MC, Dixon PM, Looney BB, Pinder JE III (1989) Estimating mean
and variance for environmental samples with below detection limit
observations. Water Res Bull 25:905-916
Okamato, K. (1991) Biological reference materials for metal speciation:
NIES fish tissue reference material for organotin compounds. In K.S.
Subramanian, G.V. Iyengar and K. Okamoto (eds.) Biological Trace
Element Research: Multidisciplinary Perspectives. American Chemical
Society Symposium Series no. 445. Washington, D.C. pp. 257-264.
Seinen, W. (1980). Immunosuppression induced by certain organotin
compounds. Vet. Sci. Commun. 3:279-287.
Snoeij, N.J., A.H. Penninks, and W. Seinen. (1987). Biological activity of
organotin compounds-an overview. Environmental Research 44:335-
353.
US Congress (1988). Organotin antifouling paint control act of 1988. HR
2210-3. One Hundredth Congress of the United States of America,
Second Session. Washington, DC.
Van den Berg, and 22 coauthors. (1998). Toxic equivalency factors (TEFs) for
PCBs, PCDDS, PCDFs for humans and wildlife. Environmental Health
Perspectives 106: 775-792.
Mackay, D., W.Y. Shiu, and K.C. Ma. (1992). Illustrated Handbook of Physical-
Chemical Properties and Environmental Fate for Organic Chemicals. Vol.
2. Lewis Pubis.
Address for TBT CRM (tissue)
Mr. Hiroyasu Ito
Chemical Division
National Institute for Environmental Studies
Japan Environment Agency
16-2 Onogawa, Tsukuba
Ibaraki 305-0053
Japan
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Comments Received from
Paul Price
March 28,2002
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Proposed Bioaccumulation
Testing Evaluation
Framework for Determining
the Suitability of Dredged
Material to be Placed at the
Historic Area Remediation
Site(HARS)
Paul S. Price M.S.
Consultant on Exposure and Risk
Issues
129 Oakhurst Rd.
Cape Elizabeth, ME 04107
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COMMENTS FROM PAUL S. PRICE (3-28-2002)
Introduction to Human Health Comments
The Remediation Material Workgroup (RMW) has charged the peer reviewers to address
a narrowly defined component of the restoration project, the establishment of HARS-
Specific Values for use in the evaluation of tissue levels in bioaccumulation studies.
As a result, of RMW discussions, specific, focused questions have been developed for
consideration by the scientific peer reviewers. These questions are directed at resolving
specific areas of concern or controversy regarding the evaluation process. However,
many of the questions are outside of my area of expertise and I have not attempted to
address them.
In these comments, I have focused on those questions that relate to the risk assessment
portion of the process. In particular, I have sought to address comments 1,18,25 and 30.
In responding to these comments I have sought to suggest alternative approaches to the
development of HARS-Specific values that 1) are internally consistent, 2) makes use of
the available data, 3) can be performed using currently available data, 4) are amenable to
refinement if additional studies of the site were performed, and 5) would allow the
quantitative description of the uncertainty in the HARS-specific values.
Response to Questions 1, 25, and 30
Three of the questions posed by the RWM deal with the design of the exposure
assessment used in the derivation of the HARS-Specific Values and the uncertainty in the
values that results form the uncertainty in the exposure assessment. Three of the
questions address the fundamental design of the exposure assessment.
1. Throughout the proposed process, there are various uncertainties introduced. Please identify the
key areas of uncertainty that need to be addressed. Are there additional data sources or
parameters that could be used to address these areas? What methods are available for describing
and accounting for these uncertainties in the calculation of HARS-Specific Values? Of the methods
available, which would you recommend for consideration and why? Please consider the
implications of implementing these methods in the regulatory framework. Please include an
evaluation of probabilistic and deterministic methods in your discussion.
25. What are the major sources of uncertainty associated with the approaches? What alternative
approaches would reduce the uncertainties? How could these uncertainties be described and
accounted for in decision-making?
30. Is the conceptual model for evaluating fish exposure to dredged material at the HARS and human
exposure through ingestion of seafood appropriate and reasonable? How can the uncertainties
associated with the assumptions in this conceptual model be reduced? Please consider the
spatial and temporal elements of exposure in your discussion.
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COMMENTS FROM PAUL S. PRICE (3-28-2002)
The approach adopted by EPA in this assessment is does not attempt to quantitatively
deal with the issue of uncertainty and relies on a deterministic approach to estimate a
reasonable maximal exposed individual (RME). This framework is then used as the basis
for establishing the HARS-Specific Values.
RME-based approaches are not amenable to quantitative uncertainty analyses since they
involve a number of qualitative or semi-quantitative determinations of whether specific
factors are relevant to an exposure assessment and what values are appropriate. The
deterministic approach is inherently a bounding approach either absolute (Theoretical
Upper Bound Estimates, TUBEs), or approximate (RME). EPA has used techniques
such as "backing off" from TUBEs (Theoretical Upper bound Estimates) or replacement of
one parameter with a typical rather than conservative value (EPA, 1992) to provide insight
into the degree of protection associated with an RME. However, this was not performed in
the assessment. As a result, no conclusion is possible on the uncertainty of the current
values. Nor is it possible to quantify the degree of protection the standards provide to any
group of New Jersey Anglers.
The RMW has specifically asked if probabilistic techniques should be used. Such tools
would be useful in the assessment fro several reasons. First, they would allow the
construction of quantitative metrics of uncertainty in the estimates. Second, they would
allow the model to better deal with variation in angler behaviors and temporal and spatial
information on contaminants in the remedial materials and in fish. Third, they will force the
EPA to more clearly define:
• The target population;
• The mechanisms of bioaccumulation;
• The exposure pathway;
• The dose determination; and
• The characteristics of the exposures.
The following is a series of recommendations for revising the assessment that make use
of such tools.
Development of a Framework for Assessing Exposure
The first step in an assessment is a clear description of the problem. This description
should include objective and testable descriptions of:
• The target population;
• The risk management goals;
• The source of exposure; and
• The exposure pathways.
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COMMENTS FROM PAUL S. PRICE (3-2&-2002)
Definition of the Target Population
It is critical that an assessment define the target population. Ill-defined populations result
in the in ability to objectively define the data necessary to assess exposure. In this case,
the population has been defined as:
A subpopulat'on of anglers that fishes exclusively at the HARS and that these anglers consume similar
quantities of recreationally caught as the Average New Jersey Angler, (pg. 20. December 21, 2001
Document)
This definition is useful in that it excludes consideration of exposure from commercial
fishing. This exclusion is a reasonable decision since commercial fish are harvested by
multiple commercial anglers who will combine fish taken from the HARS with fish
harvested from other areas. Commercial fish will be mixed with other fish as they are
sorted by species, weight, and condition, distributed through buyers, central food
distribution facilities, and taken to multiple supermarkets and fish markets. Fish that enter
into processed fish foods (fish sticks, fish cakes) are mixed even further.
During this process, it is quite likely that the fish from HARS will be diluted with catch from
other locations. As a result, individuals consuming either fresh or processed fish are likely
to only experience the intake of the occasional and isolated fish or fish meal that is
affected by HARS contaminants.
Recreational anglers are a reasonable choice for the target population. Recreational
anglers are able to harvest a large number of fish from one trip to HARS and freeze the
fish for long-term consumption. Thus, one fishing event could result in multiple exposures
over time. In addition, angler preferences in fishing location may result in repeated visits to
HARS. Therefore, this population appears to be a reasonable selection for the target
population.
The target population of recreational anglers needs to be objectively defined in terms of
actual exposure. One option is to define the population as individuals who are exposed to
HARS contamination. In this case, the definition might be, "any individual who consumes
a single meal of HARS-impacted fish at any time in his or her life as the result of
recreational angling." Such a definition would include a very large number of individuals
with trivial exposures. These individuals could dilute out the avid anglers in a probabilistic
analysis. The alternative is to define the population in terms of a specific intake of fish (7.2
g/d). The problem with this definition is that it is difficult to justify the selection of the value
and the size of the population receiving this intake may be difficult to determine.
As a result, the most appropriate approach may be to define the population in terms of
those who consume one or more meal fish affected by HARS and to be sure that the
upper bound of the distribution is well characterized.
The proposed definition of the target population; however, does not correspond with any
existing population of recreational anglers in any meaningful way. It is unlikely that there is
a group of anglers who only fishes the HARS. This definition of the target population
should be replaced with a definition that is based on the known behaviors of New Jersey
marine recreational anglers.
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COMMENTS FROM PAUL S. PRICE (3-28-2002)
Demographics of marine recreational anglers are tracked by National Marine Fisheries
Service. New Jersey was covered by the 1991 and 1990 Marine Recreational Fishery
Statistics Survey (http://www.psmfc.org/recfin/mrfssov.html Data for New Jersey anglers
is given in Table 1 and can be found in full at:
http://www.nefsc.nmfs.aov/read/socialsci/recsurvev/a7.html
Table 1. National Marine Recreational Fishery Survey - Recreational Angler Demographics for
New Jersey
Angler Category
Party/Charter
Private/Rental
Shore
Current Age
Percent
Percent
Percent
16-25
8.9
6.3
9.6
26-35
17.2
19.1
24.9
36-45
27.6
23.4
23.1
46-55
21.7
25.8
18.6
56-65
12.3
15
12.9
>66
12.3
10.4
11.1
Gender
M=89.7
M=90.4
M=91.1
F=10.3
F=9.6
F=8.9
Years Fished1
0-5
19.1
11.9
19.5
6-10
16.2
9.7
14
11-15
12.7
9.2
10.2
16-20
11.8
16.3
13.7
21-25
4.9
8.8
8.1
26-30
11.3
12
10.8
>30
24
32.1
23.8
Household Income
Less than 15,000
3.4
3
6.1
15,001-30,000
17.1
15.2
16.8
30,001-45,000
32
24.4
27.8
45,001-60,000
24
24.8
22.7
60,001-85,000
11.4
18.4
16.2
85,001-110,000
8
10
6.5
110,001-135,000
1.7
1.9
1.3
135,001-165,000
1.7
0.9
1
>165,000
0.6
1.4
1.6
First, the target population is actually two populations2 with differing potentials for
exposure. The first population is anglers on privately owned or rented fishing boats. The
1 Note that this includes fishing that occurs outside of New Jersey
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COMMENTS FROM PAUL S. PRICE (3-28-2002)
key factor here is that the angler captains the boat and determines the location for fishing.
If an owner or renter decides to favor HARS, then his or her exposure would be increased
by repeated harvesting of HARS-affected fish.
This population will be limited to boat owners, families of owners, and friends of owners.
This population will tend to be affluent individuals and their families. Such anglers would
be expected to have mobility similar to that of the general population. In addition, the age
structure will be older then the general population. Finally, consumption of fish will be
performed primarily because of personal preference for fresh fish and not for economic
reasons. Thus, these anglers are not "subsistence" anglers.
The second population is the anglers who use charter or party boats. There a number of
commercially available boats located on the North Coast of New Jersey or who service the
Bight:
http://www.tealfishinq.com/
http://www.noreast.com/charterboats/chartemewvorkbiaht.cfm
http://www.noreast.com/partvboats/BoatList.cfm?aiftsQnlv=false&area=New%20York
%20Biaht
These anglers have much less control on the location that they fish then angler on private
boats. The boat captain makes the decision on where to fish based on the season, tides,
and reports from other boats on fishing successes. Anglers using these vessels cannot
choose the location for fishing. In all likelihood, anglers in this sub population will at best
be "ocasional" anglers at the HARS. However, successful anglers would in a single trip
bring home sufficient fish for multiple meals.
Like the private boat anglers, party boat anglers would also be expected to have mobility
similar to that of the general population. In addition, their age structure is like boat owners,
older then the general population. Finally, consumption of fish will be performed primarily
because of personal preference for fresh fish and not for economic reasons (the cost of
the boat trip is typically greater than the value of the fish caught). Thus, these anglers are
not "subsistence" anglers.
Children of anglers are a population of great concern. Anglers bringing fish home would
be expected to share their catch with children. EPA needs to determine if children of
recreational anglers are included in the population. If they are included then the intake of
fish for children will need to be identified. This point is also discussed later in these
comments.
Once the target population is identified, the definition of the exposure pathway can be
related to the behaviors of the target population.
The Exposure Pathway
The exposure pathway in the assessment is the consumption of fish that have
accumulated contaminants from the HARS remedial material. This pathway can be
2 The population of anglers may also Include a third sub-population, shore anglers, if the agency determines that
fish can range from the HARS to the shore.
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COMMENTS FROM PAUL S. PRICE (>28-2002)
decomposed into three components and the uncertainty in those components can be
evaluated. The components are:
1. Relationship of the proposed standards to spatial and temporal variation in
materials at site during and following remediation.
2. Relationship between the spatial and temporal variations in concentrations of the
contaminants at site to the levels in the fish and shellfish consumed by the target
population and the temporal and spatial variation in those levels.
3. The consumption of fish affected by HARS and the resultant intake of the
contaminants in the population of concern.
The following is a discussion on each of these points.
Uncertainty and bias in the relationship of the proposed standards in tissue to
spatial and temporal variation in materials at site during and following
remediation
The proposed approach does not discuss this issue in an explicit manner. However, the
risk assessment used in setting the standards implicitly assumes that the standard for a
given contaminant will be determined based on a finding of what a safe level would be if it
occurred in the benthic organisms over the entire HARS site.
In reality, it is unlikely that all materials places at the HARS will result in tissue level that
are at or just below the HARS-specific values for all of the contaminants. Some materials
will contain low levels of ail of the contaminants, some will have a single elevated
contaminant that is just below the HARS-specific value, and spme will have a mixture
where all are below the their respective HARS-specific values.
The net result will be that upon completion of remediation the average concentration in the
benthic organism are likely to be much lower than the standards established for each
contaminant. The magnitude of this difference could be tracked by reviewing the actual
tissue levels of the contaminants of materials disposed at the site and the volumes of
those materials.
A second issue is the persistence of the elevation of tissue levels over time. It is possible
that test samples of a sediment will produce tissue levels in bioaccumulation studies may
decline over time as organic compounds disperse or degrade. EPA assumes that the
level in benthic organisms will remain elevated for 70 years. EPA should give
consideration to the potential of contaminants to disperse, degrade, or to be come less
bioavailable over time where there is data to support such changes.
Uncertainty in the relationship between the tissue concentrations of
contaminants at HARS and whole body levels in the fish and shellfish consumed
by the target population
EPA Region 2 proposes to assume that all fish consumed by the target population will
have residues that are equal to the product of the site use factor, trophic transfer factor
(TTF), and the tissue value. This assumption simplifies a very complex spatial and
temporal relationship and is likely to be highly conservative.
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COMMENTS FROM PAUL S. PRICE (3-28-2002)
Each species has a distinct relationship between to the contaminants in the remedial
materials. However, the fish (or the fish's prey) must be present at the HARS to
bioaccmulate the contaminant. The HARS a large area but is only a small fraction (5%) of
the New York Bight Apex.
EPA has assumed that if a fish species is present in the Bight then the fish will have a
tissue level that is equivalent to the tropic transfer factor times the tissue level. This will
only be true if the benthic organisms on which the species feeds or which the species prey
feeds on are within the HARS. Fish that range over an area greater than the Bight, or
who's prey range over an area greater than the Bight, cannot be at equilibrium with the
benthic organism in which they directly or indirectly feed. EPA has not considered this
factor in the proposed approach.
While I have not reviewed all of the assumptions in the manuscript "Risk- Based
Management of Contaminated Sediments: Consideration of Spatial and Temporal
Factors" (Linkov et at., 2002). The paper presents the basic concepts of the role of spatial
limits and feeding ranges and suggests a reasonable approach for modeling
bioaccumulation a specific species (flounder).
In this paper, the authors assess a species with a limited feeding range (250 ha). Had the
model been applied to piscivorous fish such as a bluefish or striped bass the range would
have been much larger. While I am not an expert in the size of the feeding ranges of fish
species, bluefish and striped bass migrate over large portions of the East Coast thus it
appears plausible that such top predators could have a feeding range the size of the entire
Bight Apex. If this were the case, blue fish and striped bass would on average be at 5% of
the level predicted by the tissue level and the trophic transfer factor.
A second issue is the spatial pattern of elevated concentrations of the contaminants in fish
that occur because of the HARS remedial materials. As the feeding range of a species
decreases the probability that the entire range will fall within the HARS increases but the
area where the affected fish can be caught decreases. In the following figure, the area
around the HARS is divided into three zones. The smallest zone (Zone A) is the HARS
and the area immediately adjacent to the HARS. Zone A is the area where flounder and
other benthic fish with small feeding ranges must be caught in order in order for an
exposure to HARS contaminants to occur. Zone B plus Zone A represents the area
where a species with a larger feeding area must be caught. Fish such as bluefish or
striped bass caught any where in Zones A, B or C (the entire Bight Apex) would result in
an exposure to HARS contaminants.
The size of the zones will be directly related to the feeding range of a species. One
approach for setting the size of the zones might be to set a species' zone based on HARS
plus the area that is within a distance D, where D is the square root of the species feeding
range.
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COMMENTS FROM PAUL S. PRICE (3-2&-2002)
HARS
Zone A
-Zone B
•Zone C
(Bight Apex)
New York Bight
This approach of defining species-specific zones offers two advantages. First, the zone
provides the spatial description for the modeling described in (Linkov et al., 2002).
Second, it defines the area where fish consumption is relevant to the establishment of the
HARS-specific values. This point is discussed in the following section.
The consumption offish affected by HARS and the resultant intake of the
contaminants in the population of concern
In the prior section, zones are established for different species of interest to recreational
anglers found at HARS. The goal of the assessment of fish consumption should be to
determine the intake of the species harvested from the affected zones. The ideal
approach would be to perform a survey of anglers at local marinas to determine factors
such as:
1. Number of trips to HARS or HARS affected zones by private, rental, charter, and
party boats;
2. Harvesting rates of fish by species in each zone;
3. Consumption and sharing behaviors;
4. Frequency of use of charter or party boats.
Surveys of angler behavior can be quite focused since ocean-going vessels depart from a
limited number of locations in the New York Bight. A second option for a survey is to place
an observer at the HARS and record the number of boats actually fishing at the HARS and
the duration of fishing activities. Such a survey would establish the frequency of anglers
who visit the site and the identification of the specific boats that visit the HARS would also
allow for follow-up interviews with anglers and captains.
In the absence of such data, EPA can use the marine portion of the existing NJMSC
survey or other relevant surveys of marine anglers. This point is discussed in detail in the
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COMMENTS FROM PAUL S. PRICE (3-28-2002)
following section. However, such data provide information on the total intake and is a
conservative estimate. EPA should seek to determine the fraction of the fishing performed
at HARS or HARS affected zones. Preliminary data of the relative popularity of the site
may be available in fishing reports for local waters see for example:
http://fishinq.iniersev.com/
The key point is that the spatial description of the contamination in the fish (zones) must
be consistently related to both the level of residue anticipated to occur in the fish and to the
amount offish harvested and consumed from that zone.
Once an angler brings a fish home, exposures can occur to a number of individuals. An
angler can consume the fish he or she catches or they can share the fish with their friends
and families. Sharing increases the size of the exposed population but reduces the
amount of fish each individual consumes. EPA may wish to consider both possibilities
when there is no data is available on how fish are shared.
Sharing also allows the consumption of fish by children. If EPA determines that different
toxicological criteria should be applied to children, then the estimate of fish intake under
the assumption of sharing (see table 2 and 3 below) can be used to assess children's
exposures.
Probabilistic models can be used to models fish consumption. As discussed below,
probabilistic models are very appropriate because of the great degree of variation in fish
ingestion rates.
Integration off the Three Steps in the Exposure Pathway
The final framework will have to consider the uncertainty that occurs for lack of data in all
three steps. Two-dimensional Monte Carlo modeling may be helpful in this process. In
fact, probabilistic models are likely to be the only way to integration the data on uncertainty
in all three steps. Probabilistic models also permit a more quantified definition of the
degree of protection that the standards are designed to achieve. For example, It is
possible to state that the fraction of the population that will be kept below the risk goals
and the confidence in that estimate.
One study that is very relevant to the HARS site is Wilson et al. (2001). In this study,
exposures to PCBs and DDT in fish that came from offshore sediments were modeled for
party boat anglers.
Greatest Source of Uncertainty
The single greatest source of uncertainty is in the linkage of the intake to the contaminant
levels in fish. If fish are at equilibrium with the benthic organism then the area of affect fish
will be largely limited to the HARS. If they are limited to the HARS they will make up a tiny
fraction of the anglers intake but the exact fraction and how this fraction is distributed
across the anglers of the Bight is unknown.
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COMMENTS FROM PAUL S. PRICE (3-2&-2002)
Estimates of Value Used for Specific Factors (Question 18)
Question 18 asks for comments on the specific values used in the assessment.
18. Please comment on each factor listed above (and in Table 5) as to its appropriateness for use in the
equations listed above. Would you recommend additional factors? Would you change or modify the
equations as written above? If so, how?
The following are comments on the values
Fish Consumption Rate
One of the most disappointing aspects of the proposed approach is the selection of a fish
consumption rate. The rate is derived from an average of angler consumption of both
fresh and saltwater fish. This approach raises a number of concerns.
First, the fish intake of freshwater angler is likely to be quite different that saltwater anglers.
Freshwater anglers have a higher percentage of catch and release anglers and many
freshwater fish are not desired for food. Fresh waters may not be as productive as marine
waters (Ebert et al. 1994). Thus using a value that is the average of fresh and salt-water
angler intake is a great concern. It is extremely unlikely that EPA's assertion that the
mean intake will apply to the subpopulation of anglers relevant to the site (see data in
Tables 2 and 3).
Second, use of the "mean intake rate" ignores the known variability in fish consumption
rates across anglers. A repeatedly demonstrated in EPA's Exposure Factor Handbook
recreational angling is a highly skewed. The top 2-5% of anglers may have intake that are
10 fold higher than the mean of the distribution. Failure to consider this variation could
result in a lack of protection of the most highly exposed anglers.
Third, the use of the mean did not take advantage of the site-relevant data contained in
the NJMSC data. The NJMSC survey collected data on fish harvesting and consumption
by region and location. Data collected included data for 31 individuals who reported
fishing and crabbing on shore or with in 10 miles of shore on the northern coast of New
Jersey (Sandy Hook to Raritan). These data are not ideal since the combing on shore
and ocean pier fishing with the near shore (<10 miles) fishing that is most relevant to the
HARS. However, the characteristics of this group are clearly more relevant to the target
population then the average of all New Jersey anglers.
Table 2 presents the estimates of consumption rates for the 31 anglers. In developing
these estimates the following assumptions were made:
• The total amount of fish caught or frozen is assumed to be the fillet weight not
whole fish;
• Where the angler reported a larger "amount frozen" than "amount consumed" the
"amount frozen" was used;
• The consumption rates for summer and fall were added together for each angler;
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COMMENTS FROM PAUL S. PRICE (3-28-2002)
• The Angler does not fish during November to June; and
• Two sets of estimates were developed one based on the assumption that the only
the angler would consume the caught fish and a second assuming that the fish
would be shared by all the members of the household.
Table 2. Fish Consumption Data for the 31 Anglers Who Fish Northern New Jersey
Coast (On Shore to 10 Miles Off Shore)
Lbs/yr
Consumption Only by Consumption by All
Angler
Household Members
Mean (All Anglers (31))
10.6
3.6
Mean (Consuming Anglers (24))
13.7
4.7
Max
62.0
15.5
g/day
Mean (All Anglers (31))
13.1
4.5
Mean (Consuming Anglers (24))
17.0
5.8
Max
77.0
19.3
Table 3 presents the same data but species not harvested at the site have been excluded
(sharks, tuna, and crabs).
Table 3. Fish Consumption Data for the 31 Anglers Who Fish Northern New Jersey
Coast (On Shore to 10 Miles Off Shore)
Lbs/yr
Consumption Only by Consumption by All
Angler
Household Members
Mean (All Anglers (31))
8.6
3.0
Mean (Consuming Anglers (21))
12.7
4.4
Max
62.0
15.5
g/day
Mean (All Anglers (31))
10.7
3.7
Mean (Consuming Anglers (21))
15.8
5.4
Max
77.0
193
Based upon this analysis, it is likely that a substantial fraction of the anglers fishing the off
shore waters near HARS may have total fish intakes that are much higher than the 7.2 g/d
value used in the analysis. These data also strongly suggest the need for approaches that
allow the consideration of variation in intake.
Finally, these data demonstrate the ability of telephone survey to collect data that are
relevant to the HARS site. EPA should consider performing a telephone-based survey of
marine recreational angler behavior in the counties closest to the HARS site.
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COMMENTS FROM PAUL S. PRICE (3-28-2002)
Duration
The assessment assumes that anglers will be exposed to the residues over a 70-year
period. The data on the demographics of New Jersey anglers does not support this value,
see Table 4.
NJ ocean angling is a recreational behavior that is taken up as an adult. A review of
duration suggests that anglers are not permanent residents (average time in the state is
seven years) and take up marine angling in middle age. This suggests that angling
durations of less than 70 would be appropriate.
Table 4. Demographic Data on 31 Anglers Who Fish Northern New Jersey Coast (On Shore to
10 Miles Off Shore)
Angler ID Number
Years in NJ
Current AGE
Age Arriving in NJ
69
7
43
36
390
5
26
21
455
6
23
17
491
8
44
36
501
5
45
40
645
8
51
43
735
8
66
58
1508
8
74
66
1656
8
37
29
1687
6
40
34
1910
8
30
22
1965
6
28
22
2001
8
55
47
2161
8
29
21
3123
8
29
21
3384
8
31
23
3752
8
31
23
4050
8
49
41
4347
7
49
42
4542
8
32
24
6045
7
45
38
6180
8
47
39
6776
8
30
22
6792
8
44
36
6972
5
36
31
7241
8
54
46
8055
8
44
36
8218
7
47
40
8330
6
36
30
8866
5
45
40
10016
6
29
23
Average
7.16
40.94
33.77
Min
5
23
17
Max
8
74
66
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COMMENTS FROM PAUL S. PRICE (3-28-2002)
Fraction of Fish Caught that are Affected by HARS
As discussed above EPA implicit assumption that all fish caught by the target population
will be affected by the HARS, is at best ill-conceived. It is highly unlikely that there exists a
population of recreational anglers who only fish the HARS site. The more reasonable
assumption is that there are populations of owners/renters of boats and individuals who
use charter/party boats who fish the site. Evidence of fishing at HARS can be seen in a
recent fishing report3. However, these anglers cannot be assumed to fish only on at the
HARS portion of the Bight.
A more appropriate model would be to assume that the anglers intake fish from both
affected and non-affected areas and that the affected areas are species specific. The
area may be the entire Bight for certain species but limited to the HARS for others.
Cooking Loss
The proposed approach does not consider the impact of cooking loss. Lipophilic
compounds such as PCBs, dioxins, DDT, and other organochlorine pesticides have been
shown to be removed during cooking because of fat loss (Sherer and Price, 1993). This
factor should be included for those chemicals.
Species Use Factors
The basis for this factor is unclear. If a fish is not present in the Bight, it will not be
consumed by recreational anglers. If it is present then it will be affected by the benthic
organisms it consumes. The more important issue is not the fraction of the year a species
is present in the Bight but the fraction of angler's intake that is affected by the HARS. This
should be addressed by species-specific zones discussed above.
Minor Comments on Other Questions
19. Are the methods used to derive the human health exposure parameters and assigned values
discussed in Section E appropriate (please review the referenced appendices)? If not, please
elaborate. How should these factors be factored into the risk analyses and decision-making?
As discussed in question 18, there are a number of modifications that would benefit the
methods for performing the exposure assessment.
3 Published in the Asbury Park Press 12/28/01
The 47-degree water pushed the last of the porgies and sea bass into the deep, but
ling have moved in to replace them on the Scotland Grounds and the Mud Buoy....
14
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COMMENTS FROM PAUL S. PRICE (3-28-2002)
20. Is the approach taken to relate fish whole body and fillet concentrations scientifically appropriate? If
not, what method would you recommend?
The method appears to be reasonable. Uncertainty in this factor appears to be small
compared to other sources of uncertainty.
21 .Could the analysis be improved by focusing on key fish (seafood) species at the HARS? What
characteristics should be used to select these key species?
Yes. The assessment should be species-specific. The species are determined by those
species which are affected by the HARS and which are consumed by recreational anglers.
22. In your opinion, is the approach for assuming total metal to be in the most toxic form appropriate and
reasonable? Should metal speqation/complexation be considered in the assessment of metals
bioaccumulation, trophic transfer, and human health risks? Is the proposed approach for evaluating
methyl mercury appropriate? Are there alternative analytical or risk assessment techniques available
that would improve the risk assessment of metals? Is the multiplier proposed for adjusting measured
concentrations of arsenic appropriate and reasonable?
Yes, it appears to be reasonable. However, EPA/ACE should investigate whether basic
considerations of the chemistry can be used to exclude metal species that are not likely to
occur. For example, Cr+6 would not be expected to occur under reducing conditions.
9. If the approach for evaluating dioxin is modified, should it include the contribution of PCBs with dioxin-
like activity as proposed? If so, how?
EPA should avoid double counting PCB toxicity in the assessment.
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COMMENTS FROM PAUL S. PRICE (3-28-2002)
References
Wilson, N.D., P.S. Price, D.J. Paustenbach. 2001. An Event-by Event Probabilistic
Methodology for Assessing the Health Risks of Persistent Chemicals in Fish: A Case
Study at the Palos Verdes Shelf, Journal of Toxicology and Environmental Health, 2001
Apr 20; 62(8): 595-642
EPA. 1992. Guidelines for Performing Exposure Assessment
Ebert, E.S., P.S. Price, and R.E. Keenan. 1994. Selection offish consumption estimates
for use in the regulatory process. J. Exp. Anal. Environ. Epid. 4(3):373-394
Sherer, R.A. and P.S. Price. 1993. The effect of cooking processes on PCB levels in
edible fish tissue. Qual. Assur. Good Practice, Reg. and Law 2(4):396-407.
16
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Comments Received from
Harlee Strauss, Ph.D.
April 1,2002
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Responses of Harlee Strauss, Ph.D. to Charge Questions
April 1,2002
Overall Process
1. Throughout the proposed process, there are various uncertainties introduced. Please
identify the key areas of uncertainty that need to be addressed. Are there additional
data sources or parameters that could be used to address these areas? What methods
are available for describing and accounting for these uncertainties in the calculation
of HARS-Specific Values? Of the methods available, which would you recommend for
consideration and why? Please consider the implications of implementing these
methods in the regulatory framework. Please include an evaluation ofprobabilistic
and deterministic methods in your discussion.
While there are many uncertainties in the various parameters, the overarching uncertainty
is the objective of the receptor to be protected. It is a recreational angler, but it is
completely unclear how conservative (protective) the analysis is intended to be. Is it a
high-end consumer who angles at the HARS? Also, is it only the current population who
angles or also potential future users? The selection of the percent of population to be
protected is especially important in light of the relatively "non-conservative" policy
decision to use 10E-4 cancer risk as the basis for calculating HARS specific values.
The TEF, as it stands, does not address either variability or uncertainty in a clear and
explicit manner. It does not address the variability of the fish consuming population,
including the high-end consumer (ingestion rate) and/or those who may eat more than the
fillet. It does not address the variability in exposure duration among the angling
population and the variability of many of the parameters that lead to the prediction of
contaminant concentrations in the finfish. Each of these parameters also has uncertainty
associated with it. Some of the specific parameter values utilized in the calculation are
quite conservative, others are central tendency, and still others will lead to underestimates
of the risk. However, there is no overarching summary of this as to the likelihood that
the proposed values are protective of human health.
The question of uncertainty and variability should both be addressed explicitly. This can
be done in the deterministic risk assessment framework utilized here by developing a
summary table and showing where the selected values fall within the ranges. This
question could also be evaluated quantitatively using probabilistic risk assessment
techniques. I recommend that both a deterministic point estimate for an explicitly defined
RME be determined and a probabilistic risk assessment be used to determine where the
deterministic RME falls within the risk range.
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2. Is the measurement of 26 priority pollutant PAHs sufficient for characterizing the
risks associated with the total PAH bioaccumulated by organisms exposed to
dredged material proposedfor placement at the HARS? Does measurement of the
alkylated compounds significantly improve risk assessment of PAHs?
The current practice in human health risk assessment does not consider alkylated PAHs
and no toxicity factors (for either cancer or non-cancer effects) are available. However, I
believe that current practice for evaluating risk associated with PAHs is inadequate, and
that including alkylated PAHs will lead to improved risk assessment. At a minimum,
measuring the presence of alkylated PAHs compounds would provide a better
understanding of one aspect of the uncertainty associated with the risk assessment, that is,
how much PAH is actually present in organisms.
Toxicological data on alkylated PAHs are sparse. However, the data that do exist suggest
that alkylated PAHs are at least as toxic as their non-alkylated counterparts. This is
clearly true of 7,12 dimethylbenzanthracene, which is classically used as a tumor
initiator. Recent in vitro studies on mechanisms for tumor promotion (specifically
inhibition of gap junctional intercellular communication) have also shown that alkylated
phenanthracenes, fluorenes, and fluoranthracenes are at least as potent as their non-
alkylated parents. 1 methyl-naphthalene is clearly more potent than either naphthalene or
2-methylnaphthalene. (Weis et al. 1998. Environ Health Perspect. 106(l):17-22; Upham
et al. 1998. EHP 106 Suppl 4:975-981). A study of the mutagenicity (Ames test) of
methyl fluorenes found that they are mutagenic if the methyl group is in a certain
position. Fluorene itself is not mutagenic in Ames assays. Enhanced mutagenicity has
also been observed for benzofluorenes, i.e., the methylated compounds are more
mutagenic than the parent (Lavoie et al. 1989. Mutation Res 91:167-76). Similarly,
Ames test mutagenicity studies of dibenzothiophene (DBT) and several alkylated
derivatives indicate that while DBT itself is not mutagenic, several alkylated derivatives
are mutagenic following metabolic activation (McFall et al. 1984. Mutation Res.l35:97-
103).
Weis et al. 1998 and Upham et al. 1998 cite studies that indicate that bioalkylation
converts noncarcinogenic PAHs to carcinogenic PAHs. This observation should also cast
a long shadow of doubt over the adequacy of current risk assessment procedures for
evaluating PAHs.
In addition to the toxicologic literature suggesting adverse effects (related to
carcinogenicity in the above cited references) of alkylated PAHs, other data indicate that
alkylated PAHs generally persist longer in the environment than their parent compounds
and tend to bioaccumulate to a greater degree (Irwin et al. 1997. National Park Service).
The combined effect of these sets of observations is that alkylated PAHs could be the
major contributors to toxic effects observed with mixtures of PAHs.
3. Is the proposed adaptation of EPA Method 8270 acceptable and appropriate for
regulatory decision-making?
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I don't have any expertise with various analytical methods. However, the analytical
method chosen must routinely meet data quality objectives including identifying the
specified alkyl substituted PAHs and quantifying them at the designated limit of
detection.
4. Under what specific conditions would the testing for alkylated PAHs for a
particular project be appropriate and warranted?
One of the objectives of the testing protocol should be to make sure that dredged material
with substantial oil contamination does not receive approval for placement at the HARS.
Non-alkylated PAHs may be associated with oil contamination, but including the more
persistent alkylated PAHs will provide a fuller picture of the contamination. Thus, it
seems prudent to always test for alkylated PAHs, and to develop a total concentration that
should not be exceeded.
5. What uncertainties would be introduced within the analysis of risk should
alkylated PAHs be included? What steps could be taken to account for these
uncertainties in decision-making? Given the likelihood the method for using non-
detects (as described in EPA/CENAN(1997) will result in an overestimate of risk,
what are the implications?
Including alkylated PAHs will introduce many uncertainties if they are treated as
individual compounds and the risk from each of them is summed. For starters, there are
no toxicity values associated with alkylated PAHs. A later charge question suggests
using the toxicity values for the parent compound. This could be a starting point,
although in some (many?) cases this may underestimate the risk as the data cited in
response 2. show that methylated forms are more toxic than the parent. However, use of
this rather crude approach may provide a more accurate picture of PAH risk than to
continue the current human health risk assessment approach of ignoring the issue of
alkylated PAHs altogether.
That said, it might be better to take a different approach at this point in our knowledge
base. As suggested in a previous response, one could develop a benchmark number for
total PAHs (linear and alkylated) based on risk. This value could be based on toxicity
tests (e.g., mutagenicity and/or in vitro tests for other endpoints) of PAH mixtures found
in dredged materials. The benchmark could be developed by using "generic" petroleum
residues in a series of toxicity tests. The guidance could allow the application of the
generic benchmark to be superceded by actually testing of the dredged material to be
disposed. The use of in vitro toxicity tests as the regulatory endpoint avoids the problem
of potentially overestimating risk because many of the individual analytes are below the
limit of detection.
6. It is recognized that additional methods have been used for the analysis of
organotins. Will the proposed analytical method (Rice et al. 19987) provide
adequate data of sufficient quality to assess relevant risks from organotins? Is this
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method the most appropriate analytical method for organotins? If not, please
provide recommendations.
I have no particular expertise in chemical analysis. I defer to the judgement of others on
the peer review panel.
7. What special QA/QCprocedures should be implemented to ensure the quality and
usability of the organotin data?
Again, I have no particular expertise in this area. I defer to the judgment of others on the
peer review panel.
8. Under what specific conditions would the testing for organotins for a particular
project be appropriate and warranted?
My understanding is that organotins are widely used in marine paints as an antifouling
agent. Thus, the organotins may be in any dredged material, and should be routinely
included in the testing and evaluation program.
9. If the approach for evaluating dioxin is modified, should it include the
contribution of PCBs with dioxin-like activity as proposed? If so, how?
The method used at the HARS should be consistent with overall agency policy. Current
EPA guidance (per the IRIS profile for PCBs) recommends subtracting the
concentrations of PCBs with dioxin-like activity from the total concentration of PCBs
and then evaluating the adjusted concentration of PCBs using the appropriate PCB slope
factor. This subtraction is intended to avoid "double counting" the carcinogenic activity
of certain congeners as both PCB and dioxin-like carcinogens. As part of the dioxin
reassessment, EPA is reportedly developing new guidance on evaluating PCBs with
dioxin-like activity.
There are additional complexities in the evaluation of the risk from PCBs that are not
considered in the current testing and evaluation framework. For example, the cancer
potency factor for PCBs recommended in IRIS and proposed for use in this testing
evaluation framework is based on animal cancer bioassays using PCBs as manufactured.
However, the mixture of PCBs in dredged material will be different from the original
mixture due to differential fate and transport properties of the individual congeners as
well as the likely blending of many individual sources. Moreover, the congener
composition (e.g., the weight percent of each congener present) of the benthic
invertebrates will differ from the congener composition of the sediment because of
differential bioaccumulation properties of individual congeners. Similarly, the congener
composition of the predator species eventually ingested by humans will differ from that
of the benthic invertebrates. Dioxin-like PCBs will likely have a higher weight percent in
food species than in the original sediment and in tested invertebrates. This uncertainty
would lead to underestimation of the risk to humans even if dioxin-like PCBs are
separately characterized in the tested invertebrates. Moreover, the carcinogenic potential
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of the non-dioxin like PCBs in the prey species will differ from the test species, although
it is not clear whether this will lead to an under or over estimate of the risk.
10. Please consider the policy for assigning values to tissue residues that are
reported as "< detection limit. " Please comment on the effects of this policy on
the outlined evaluations. If the current approach is considered inappropriate,
what would be a more technically supportable alternative.
The current approach for analytes with detection limits reported at or below the MDL is
consistent with the approach commonly taken in human health risk assessments. The
approach for samples with high detection limits is different from that commonly used.
Often, in human health risk assessments, one uses the same approach no matter what the
detection limit. This may be a better approach in some circumstances. For example,
many times the reason for the elevated detection limit is the need to dilute the sample
because of the high concentration of another analyte. If the concentration of the high
concentration analyte is expected to correlate with analytes with NDs with high DLs, then
continuing to use lA DL is appropriate. However, if there is no reason to expect the
concentrations to covary, V2 the DL (and especially the zero to full DL approach in the
EPA/CENAN policy) is likely to overpredict concentration and risk. From a risk
characterization point of view, a reasonable way to proceed would be to try to separate
the high concentration analyte from the rest of the sample and then determine the true
concentration (or at least the ND with lower detection limits) of the other analytes. While
this approach is more costly analytically, it could be cost effective if it changes the
decision regarding whether material can be classified as remediation material.
11. Is the use of functional groupings in statistical comparisons to reference
appropriate and/or preferable to statistical comparisons using individual
contaminants for the purposes of risk analysis?
The use of the functional grouping described in the text (total PCBs, total DDT, total
endosulfans, and total BaP equivalents) can be appropriate in some circumstances. Three
circumstances where functional grouping may be appropriate are:
• Interconversion (or perhaps conversion as in a degradation pathway) among the
analytes in the environment. DDT,DDE and DDD and the endosulfans meet this
circumstance
• Similarity of mechanism and toxic endpoint. This is the underlying rationale for
toxic equivalency factors, such as those applied to PAHs and to dioxin-like
compounds for carcinogenicity evaluations.
• Analytes that always occur as part of a mixture such as PCBs and PAHs
However, one has to be careful when grouping especially when modified by toxic
equivalence factors. For example, while PAHs may be appropriately evaluated for
carcinogenicity as BaP equivalents (or perhaps not, but let's say you can for the sake of
argument), this is NOT the appropriate metric for noncancer effects.
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Where grouping is appropriate for risk characterization, it is also acceptable for making
statistical comparisons between organisms subjected to test and reference materials.
Adjustment to Steady State: Organic compounds
12. Is it appropriate to apply a multiplier based on log Kowfor these compounds
(organics), or are there other specific data that can be used to estimate steady
state? If so, please identify.
I have no particular knowledge in this area and defer to the judgement of other peer
reviewers.
13. Given the increased hydrophobicity of alkylated PAHs, is the use of the correction
factor associated with the corresponding parent an appropriate approach for
estimating steady state residues of alkylated PAHs? If not, please elaborate.
If log Kows form the basis for the multiplier, and a Kow can be calculated (or has been
measured) for an alkylated PAH, then the Kow specific for the alkylated PAH should be
used. There is no reason to use the parent PAH default if a more specific value can be
obtained for the alkylated PAH.
14. For the DDT derivatives and dieldrin, please comment on the appropriateness of
using M. nasuta data rather than N. virens-specific data in the estimation of
steady state multipliers.
I note that for PCBs, where both sets of data are available, N. virens shows a slower
approach to steady state than M. nasuta, resulting in a larger multiplier. It is not clear
from the data presented whether this reflects variability among individual organisms or
real differences between the two species. A more in depth discussion of the underlying
data would be helpful to understanding this. If there is a real between the two species,
then perhaps there should be a correction to account for this.
15. Are the approaches taken to adjust organic contaminant bioaccumulation data to
steady state adequate? Do the proposed multipliers agree with previously
published studies (i.e., do they appear reasonable)? If not, please elaborate.
It is appropriate to adjust to steady state. I have no particular knowledge as to whether or
not this approach is adequate and defer to the judgement of other peer reviewers on this
question.
16. What are the major sources of uncertainty associated with the approaches? What
alternative approaches would reduce the uncertainties? How could these
uncertainties be described and accounted for in decision-making?
I usually prefer to use data from measurements rather than estimate on the basis of Kows
or other physical chemical parameters. Thus, if more data were available on longer-term
exposures compared to 28-day exposures, this would reduce the uncertainty.
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Adjustment to Steady State: Metals
17. In your opinion, is the methodology followed to derive the steady state multiplier
for non-essential metals (i.e., a factor of three) scientifically appropriate
(Appendix G)? Please elaborate. Do you have any recommendations of
additional or alternate methodologies or information that can be used to either
supplement or replace the proposed method?
It seems to me that what is being measured here is variability in the concentrations of
several metals in polychaete tissues. On visual inspection, there doesn't seem to be a
correlation between tissue concentration and sediment concentration for any of the
metals. For example the highest tissue concentration does not correspond to the highest
sediment concentration, etc. Therefore, calling it a steady state adjustment doesn't make
any sense. It would make more sense to explicitly account for variability in the risk
assessment.
Human Health Evaluations: Overall
18. Please comment on each factor listed above (and in Table 5) as to its
appropriateness for use in the equations listed above. Would you recommend
additional factors? Would you change or modify the equations as written above?
If so, how?
The equation is based on a deterministic calculation of the risk. In view of the inherent
variability of some of the parameters along with uncertainty associated with all of the
parameters, the derivation of the HARS specific values would benefit from using a
probabilistic approach that separately evaluates variability and uncertainty. The
economic impact of the ultimate HARS specific values is likely to justify the additional
cost of using a probabilistic approach.
The ingestion rate and other factors in the equation assume that only finfish are caught
recreationally and that only fillets are consumed. Neither of these assumptions is likely
to be true. First, the HARS could be used for harvesting shellfish. If shellfish are
harvested recreationally (in the future), this will affect the ingestion rate, assumptions on
trophic transfer, and the whole body to fillet ratio. Second, the equation assumes that
people ingest only fillets. However, the NJMSC (1994) data show only 51% of fish
consumers always use fillet fish with trimmed skin and fat. Half the population at least
occasionally uses whole, gutted fish and 5% reuse the fat or oil from cooking fish.
Certain subpopulations use the entire fish for cooking and making sauces and pastes.
Thus, in many cases, the whole body to fillet ratio is not appropriate. The equation does
not include a term for cooking loss prior to ingestion. While it may be appropriate to set
cooking loss to zero for certain (if not all) of the contaminants, this needs to be discussed
and justified in the text.
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The text states that recreational anglers represent an RME for assessing risks to humans,
specifically, anglers who fish exclusively at the HARS. I assume that this means that of
the total population who may ingest seafood from the HARS, the recreational angler will
ingest more than those who purchase fish or shellfish at a market where commercially
caught fish, some of which may have been caught at the HARS, are sold. However, the
analysis must be more specific than this. Is it intended to protect the high-end
recreational angler or a "typical" recreational angler? The current analysis addresses this
issue by assuming an angler who fishes only at the HARS, but ingests finfish fillets at the
rate of a typical New Jersey recreational angler. I don't find this approach very satisfying,
as fishing only at the HARS is almost certainly unrealistic and will lead to a substantial
overestimate of the risk. To balance this, typical ingestion rates will not protect high-end
recreational consumers and those who ingest more than the finfish fillets. While this
overestimate/underestimate approach may balance out, a better approach for a
deterministic calculation would be to get a reasonable estimate of the percentage of
HARS use and incorporate an ingestion rate based on a high end consumer. A still better
approach would be to develop a probabilistic analysis that incorporates the variability of
fishing locations and ingestion rates.
In summary, I recommend two different equations be used, one based on a deterministic
approach and a second based on a probabilistic approach. For the deterministic equation,
I recommend the following changes:
1) Add a cooking loss term (which may be zero for an RME, but still should be
included),
2) Add a fraction of (marine) fish caught at the HARS site and modify the ingestion rate
to reflect total recreationally caught (marine) fish, and
3) Reconsider the use of the fillet to whole fish parameter after rethinking the population
to be protected now and in the future (e.g., How frequently is whole fish prepared and
do you want to protect folks doing this? Should shellfish ingestion be included).
19. Are the methods used to derive the human health exposure parameters and
assigned values discussed in Section E appropriate (please review the referenced
appendices)? If not, please elaborate. How should these factors be factored into
the risk analyses and decision-making?
The parameter values used, as listed in Table 5 and the text, are actually a mixture of
central tendency and upper bound values. Some discussion of where the point values lie
in the overall distribution should be included. For example, the assumption of a 70
duration of angling and consuming recreationally caught fish, all of which occurs only at
the HARS, is a very conservative assumption. However, assuming ingestion of only
finfish fillets is a central tendency assumption, at best. I recommend that the objective of
who is being protected by the HARS values be clearly spelled out, and the
"conservativeness" of the individual assumptions be addressed specifically. The latter
point could also be addressed by using a probabilistic risk assessment.
A detailed review of the individual parameters in Table 5 follows.
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Cancer potency factors and reference doses
The values assigned to these factors appear to come from the most recent version of IRIS
and other well reviewed EPA sources. This is appropriate. However, the units are not
listed in the table and much confusion could occur. The table provides values for the
cancer potency factor in (mg/kg-day)-l and for the reference dose in ug/kg-day. They
should both be in milligrams. The RfD units are particularly problematic in that eqn 1
assumes the reference dose is in mg/kg-day units. N.B., the noncancer HARS specific
concentrations listed in Appendix F are too high by a factor of 1000 because of this error.
Trophic transfer
I have no particular expertise in trophic transfer, although I do have a few comments.
Generally, the methodology seems fine, although some commentary in the human health
section on whether or not the factors are considered conservative or best estimate would
be appropriate. The low trophic transfer for PAHs is consistent with their metabolism in
finflsh. I would be more comfortable if there were data for alkyl PAHs showing that the
trophic transfer of the alkyl PAHs was similar to the parent compound.
The adjustment factors to account for differences in trophic transfer of individual
congeners in PCB mixtures can lead to problems, although I don't have a better solution
than provided in the TEF framework. The main problem is that the CPF or RfD with
which the PCB toxicity is being evaluated is based on the original composition of an
Arochlor. Because of differences in trophic transfer rates (not all of which are due to
Kow, some may be due to steric factors or competing fate processes) the PCB mixture in
the edible tissues of the fish are likely to have different toxic potencies than the PCB
mixture for which we have quantitative toxicity information.
Whole Bodv:Fillet Ratio
The TEF, as proposed, uses a whole body:fillet ratio of 1.35 for all organic compounds.
This value is based on data that are not presented in the document or appendices. The
data from Bevelhimer et al. (1997), provided as the reference for determining the
inorganic ratios, also has data for organics. These data show different relationships for
PCBs, chlordane and total DDT although these data are all for freshwater species.
As stated previously, I do not think it appropriate to always convert (reduce for the
organochlorines and PAHs) the whole body data to fillets because not all fish consumers
always eat fillets. Nor is the use of this ratio appropriate if seafood other than finfish are
ingested. However, even if you assume only fillets are ingested and a conversion
appropriate, a fixed value of 1.35 has not been adequately justified in the document.
There will almost certainly be variability among the organics and among fish due to
different levels of fat, and to different lipid content of the fillet (indeed, the latter was
included as a variable in several of the conversion equations provided by Bevilhimer et
al.). No mention is made regarding the fish species that form the basis of 1.35 and a
comparison with fish likely to be consumed from the HARS, and how the ratio may be
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similar or differ due to lipid characteristics. In addition, there is no discussion of the
variability in the data and where the value selected falls in the distribution. Finally, the
use of two decimal places in the number conveys an incorrect confidence level in the
value.
For several inorganics (Cd, Cu, Pb, Ni, Zn), the TEF document incorporates a whole
bodyrfillet ratio of 1. Functionally, this means that the ratio is not considered. This is
appropriate in that Bevilhimer et al. (1997) data indicated no relationship between the
concentration in the fillet and the whole body concentration, which was invariant at the
low contaminant concentrations in their fish samples. The ratio for mercury of less than
one is appropriate as several references indicate higher mercury concentrations in the
fillet than the whole fish. However, it should be emphasized that the dataset is small, the
fish are freshwater, the range of contaminant concentrations is small, and for many, the
concentrations are near background and/or the limit of detection.
Site Use Factor
The site use factor, which is used to adjust projected fish tissue concentrations for the
fraction of time they spend at the HARS, is a very crude measure. It seems that it could
be fine-tuned by considering foraging area and known migratory patterns. This is also a
factor that would greatly benefit from targeting a few key species and developing
parameters appropriate to them.
Exposure Duration of 70 years
The TEF, as drafted, uses an exposure duration of 70 years. In other words, it is assumed
that HARS caught fish is ingested every year for a 70 year lifetime. Exposure duration is
a parameter with inherent variability; different members of the angler population will eat
fish over different lengths of time. The 70 year parameter value is clearly at the high end
of the distribution, although there may be some lifelong anglers who actually do fish at
the HARS for that length of time.
For an angling scenario such as the HARS, it is appropriate not to use EPA default values
for the length of residency in a single house (RME 30 years), as people move within a
region and still fish at the same location. Angler surveys in other areas show that there
are lifelong residents of the area who have fished for most of their lives, although they
may not have eaten the fish regularly when they were young children. The report should
provide more justification for the selection of the 70 year value in the deterministic
calculation and discuss its implication in the extent of health protectiveness of the final
HARS specific value.
Seafood Ingestion Rate (IR) of 7.2 grams/dav
The IR parameter value has a very high level of both variability (i.e., it will differ among
individuals who consume recreationally caught seafood) and uncertainty (i.e., the dataset
for estimating the ingestion rate is weak and incorporates many assumptions that are also
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weakly supported). Overall, based on a review of the NJMSC 1994 data that was used as
the source of this calculation, the 7.2 g/day figure is probably too low to use in a point
estimate calculation.
The underlying dataset for calculating the ingestion rate is a fish consumption survey of
New Jersey Anglers prepared by the New Jersey Marine Sciences Consortium (1994).
As pointed out in Appendix I, this study uses a survey design that asks about seafood
consumption in the week preceding the survey (October-November). This limits the
results in two ways. First, it is difficult to extrapolate from a one week to an annual
pattern. Second, October-November represents only one fishing season, and one of the
least active quarters for the two most highly favored species recreational species: fluke
(summer flounder) and bluefish (Appendix J and NJMSC 1994). Additionally, the
survey data themselves show that anglers fish in saltwater approximately twice the
number of days in the summer than the autumn. The first limit (the weekly to annual
extrapolation) may tend to bias the results high although the data and discussion in the
report suggest that the short term data should be representative of the longer term. The
second limit will almost certainly bias the results low.
The amount of reported fish ingestion was calculated in two ways. In the most
straightforward, the annual estimate of recreationally caught fish was directly converted
to a daily ingestion rate. This resulted in the 7.2 grams/day value. The other estimate,
based on the same survey, was based on the species-specific data in the survey. The
ingestion of all recreationally caught saltwater finfish that could be at the HARS were
summed up. This rate was then reduced by 40% to account for the percentage of
potentially HARS related fish that are not home-prepared and thus assumed not to have
been recreationally caught at the HARS. This second calculation leads to a 6.8 gram/day
ingestion rate. As a minor point, the high uncertainty in the data and the roughness of the
calculation do not justify presenting the results to 2 significant figures. Both calculation
methods yield an IR of 7 grams per day.
An alternate approach to calculating the ingestion rate, based on the NJMSC (1994)
report, would be to use the data on catch kept for consumption. The report states that an
average of 9 lbs. of flounder and fluke per catch is kept, 7.6 lbs. bluefish and 7.5 lbs. of
crab (p. 4-64). These data (at least for fluke and bluefish) could be used, along with an
estimate of fishing frequency for each species and number of persons sharing the catch,
in an estimate of ingestion rate. A correction for edible portion may also be needed for
this calculation.
Yet another approach to calculating ingestion rates could also be taken from the NJMSC
(1994) report based on the survey data showing that those who eat recreationally caught
fish ingest fish 2.5 times/week, with a salt water fish portion size of 7.3 oz (p. 4-48).
This leads to an ingestion rate of 74 grams/day. This rate, while very high, is not
completely out of line with 95th percentile estimates of fish ingestion rates for freshwater
recreational anglers (Hudson River risk assessment 32 g/d; Fox River risk assessment 59
g/d). If this approach were taken for ingestion rate, a factor for the fraction of fish caught
at the HARS would have to be included in the equation.
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Appendix I points out that lobster consumption is not included in the ingestion rate as
there is no directed recreational lobster fishery at the HARS. This raises a troubling
problem: the ingestion rate is based entirely on current conditions and does not contain
any allowance for increased use of the HARS in the future. In my opinion, the ingestion
rate should allow for realistic future use of the HARS based on improved conditions
there. Explicit consideration of future use may or may not change the ingestion rate, but
it should be included in the evaluation.
20. Is the approach taken to relate fish whole body and fillet concentrations
scientifically appropriate? If not, what method would you recommend?
See response to question 19.
21. Could the analysis be improved by focusing on key fish (seafood) species at the
HARS? What characteristics should be used to select these key species?
I do think the analysis could be improved by focussing on key species. The primary
selection criterion should be species targeted by recreational anglers at the HARS for
consumption. A second criterion could be species that bioaccumulate contaminants to a
high extent (e.g., high fat, low metabolism of compounds such as PAHs) for which there
is also consumption. Lobsters and crabs should be considered if there is a potential for
recreational harvesting of these favored seafood items in the future.
Based on my reading of the fish consumption survey conducted by NJMSC (1994) only a
few species dominate the recreational harvest, namely fluke and bluefish A more
focussed evaluation of these species will allow better estimates of consumption rates,
trophic transfer, and time spent in the HARS area.
22. In your opinion, is the approach for assuming total metal to be in the most toxic
form appropriate and reasonable? Should metal speciation/complexation be
considered in the assessment of metals bioaccumulation, trophic transfer, and
human health risks? Is the proposed approach for evaluating methyl mercury
appropriate? Are there alternative analytical or risk assessment techniques
available that would improve the risk assessment of metals? Is the multiplier
proposed for adjusting measured concentrations of arsenic appropriate and
reasonable?
The assumption that total metals are in the most toxic form is appropriate and reasonable.
The TEF allows the applicant to develop speciated data for dredged materials that fail a
total metals screening-level evaluation. This provision is appropriate for metals not
likely to interconvert between forms under environmental conditions. I have reservations
regarding its appropriateness for mercury as inorganic mercury can be biotransformed to
methylmercury in the sediment (Water Quality Criterion for the Protection of Human
Health, Methylmercury, Jan. 2001. EPA-823-R-01-001), unless there is some
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demonstration that there is a maximal extent of conversion (less than the total) due to
equilibrium or other conditions.
23. Is the assumption that the potency of alkylated PAHs can be estimated by the
potency of the parent PAH appropriate? Is this assumption likely to result in an
under- or overestimate of the risk associated with the alkylated PAHs?
See responses to questions 2 and 5. In brief, based on the sparse data on mutagenicity, it
appears that using this approach for alkylated PAHs will result in an underestimate of the
mutagenicity (and by implication carcinogenicity), but the underestimate will be less than
the current approach of ignoring the alkylated PAHs altogether.
24. Please comment on the potential for human exposure to PAHs through
consumption of finfish and other seafood.
In finfish, human exposure to PAHs will primarily be to products of metabolism rather
than the parent compound. In general these concentrations will be low, as the purpose of
the metabolism is to ready them for excretion. However, most of the data regarding
metabolism is for non-alkylated PAHs. It is not clear to me what the relative rates of
metabolism would be for alkylated/non-alkylated PAHs with the same ring structure.
It is my understanding that lobsters and crabs do not metabolize PAHs in this way, and
thus there is higher potential for PAH exposure from consumption of these seafood items.
25. What are the major sources of uncertainty associated with the approaches
described in Section E? What alternative approaches would reduce the
uncertainties? How could these uncertainties be described and accounted for in
decision-making?
This question has largely been addressed in previous comments. In summary, each of the
individual parameters has some level of uncertainty associated with it. These are poorly
discussed and described in the text at present, nor is there a clear summary whether
individual parameter values should be viewed as central tendency, RME, or other
estimates, and where the result is in after all the individual parameters are aggregated
together. This could and should be addressed by further clarity in the descriptions of
parameter values in the deterministic assessment. It could also be appropriately
addressed in a probabilistic assessment. A decision-maker should know whether the
HARS specific values are protective of a typical recreational angler or a high-end
recreational angler. The decision-maker should also know whether these values would be
protective of future use of the HARS if the future use is more intensive. Finally, the
decision-maker should be reminded that the values were developed based on an excess
lifetime cancer risk of 10*4, the riskiest end of the US EPA risk range. A probabilistic
assessment will help inform a decision-maker whether the value will protect a typical or
high-end recreational angler (variability) and with some understanding of the degree of
uncertainty.
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26. What is your recommendation for evaluating the potential toxicity of organotins?
Should they be evaluated as individual compounds? Summed as total? Should
there be some consideration of relative toxicity?
It is my understanding that the various organotin compounds can interconvert among
each other. If this is the case, then the total organotin concentration should be used in the
toxicity assessment.
27. Please comment on the appropriateness of the proposed approach for converting
and using the analytical data for alkylated and parent PAHs to estimate riskfrom
all PAHs.
See 2,5, and 23.
Human Health Evaluations: Comparison to HARS-Specific Values
28. Do you believe that the "disaggregate " modeling discussed above (and shown in
Figure 4) for estimating human health HARS-Specific Values for lead is
appropriate? Would you recommend an alternative risk assessment method be
used given the information and data available? Do you believe the method
described has appropriately taken uncertainty into account? Please elaborate.
Lead is commonly handled differently from other contaminants in risk assessment since
we know so much more about lead toxicity and its low dose effects compared to most
other toxicants in question. On balance, I think this is appropriate, although it does have
the disadvantage of not considering potential interactive effects of lead with other
developmental toxicants such as PCBs and methylmercury.
While it is not clear from the write up in the TEF, I think the lead model for children was
used in the evaluation. However, the fish consumption evaluation (e.g., ingestion rate,
body weight, etc) is based on adults. This disconnect should be addressed, either by
using a (lower) fish consumption rate for children or by using an adult lead model (which
will result in higher allowable lead levels).
Human Health Evaluations: Consideration of Combined effects
29. In your opinion, are the methodologies and equations described above
appropriate for estimating total carcinogenicity and combined non-cancer
impacts of contaminant mixtures accumulated from dredged materials proposed
for use as Remediation Material at the HARS?
The equations and methodologies for characterizing the total carcinogenicity of the
contaminant mixtures in the proposed dredged materials are consistent with widely
applied and accepted risk assessment methodology. In my opinion, it is suitable in this
situation as well. The non-cancer impacts should be evaluated for all contaminants, not
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just those listed in the Table. Carcinogenic compounds also have non-carcinogenic
effects, and should be included in the evaluation of total non-cancer effects.
The overall approach for evaluating the combined effects of multiple contaminants, i.e.,
evaluating the individual contaminants at a specified risk level, and then evaluating the
combination for the same risk level, differs from that used in some other EPA programs.
For example, for the preliminary remediation goals used in Superfund, a lower risk level
(frequently 10-fold) is used for the individual compounds than what is intended for the
total site risk. However, the approach for combining effects proposed for the TEF should
be fully protective and allows for a large amount of variability among different dredged
materials. The proposed TEF approach allows for materials with relatively high
concentrations of a single contaminant or relatively lower concentrations of multiple
contaminants to be evaluated against the same risk standard.
30. Is the conceptual model for evaluating fish exposure to dredged material at the
HARS and human exposure through ingestion of seafood appropriate and
reasonable? How can the uncertainties associated with the assumptions in this
conceptual model be reduced? Please consider the spatial and temporal elements
of exposure in your discussion.
The receptor that forms the basis for the HARS specific values has the following
characteristics:
A 70 kg adult angler who ingests 12 meals/year (8 oz/meal) of fish fillets caught at the
HARS every year for 70 years. The fillets are cooked in such a way that there is no loss
of contaminants through cooking.
Other key components of the conceptual model are:
The fish consumed spend % of their lives at the HARS where they are exposed to the
contaminants through the foodchain only. Trophic transfer can be modeled using the
Gobas model and a contaminant specific trophic transfer rate with a 28 day
bioaccumulation test, corrected for steady state, as the basis for contaminant
concentrations in the prey items.
As stated in response to previous questions, I think the definition of the receptor needs
work, including clarification of who is to be protected. Once defined, both variability
and uncertainty needs to be explicitly considered in the risk assessment. This should be
done both qualitatively within the context of a deterministic assessment and
quantitatively within the context of a probabilistic assessment.
Based on the current definition of the receptor alone, it appears that the HARS specific
values may not be protective of a high-end consumer. The data provided in NJMSC
(1994) indicates that one fishing trip to the HARS will result in adequate fish to prepare
the number of meals assumed for the year, and a significant proportion of the population
will prepare whole, gutted fish, not just fillets. Some of the population (5%) will use also
consume the oil drippings, clearly showing that for some, no cooking loss is an
appropriate assumption and considering fillets only is underprotective. While the 70 year
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exposure duration may be too long on the average, studies of recreational anglers in other
areas indicate that there are lifelong fishermen. On balance, this exposure scenario is
likely to protect a typical angler, but not a frequent recreational angler from cancer risk,
which assumes a 70 year exposure. It is likely (but should be better shown) to be
protective of recreational anglers from risks of non-cancer effects.
The degree of conservativeness of the estimate of contaminant concentrations in the fish
due to concentrations in the sediment is the also key to determining whether the HARS
values will protect a frequent recreational angler. Many variables and assumptions are
involved in the fish contaminant concentration calculation. It would be helpful to target
specific species that are consumed by anglers to reduce the number of assumptions
necessary to include many fish types. For example, if bluefish and fluke were the focus,
one could get better estimates of the time they spend at the HARS based on their normal
foraging range and the area of the HARS. More explicit data could also be used for fat
content (or variability of fat content) of the fish for trophic transfer rates.
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Comments Received from
Rick Wenning
March 25,2002
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DRAFT
SCIENTIFIC PEER- REVIEW
"Proposed Bioaccumulation Testing Evaluation Framework
for Assessing the Suitability of Dredged Material to be Placed
at the Historic Area Remediation Site (HARS),
DRAFT December 21,2001"
Consideration of Human Health Issues
Prepared for:
United States Environmental Protection Agency
Region 2
New York City, NY.
Prepared by:
Richard J. Wenning
ENVIRON International Corporation, Inc.
riwenning@.environcorp.com
Emeryville, California
DRAFT March 25,2002
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SCIENTIFIC PEER- REVIEW
"Proposed Bioaccumulation Testing Evaluation Framework for Assessing the
Suitability of Dredged Material to be Placed at the Historic Area Remediation Site
(HARS), DRAFT December 21,2001"
TABLE OF CONTENTS
I. EXECUTIVE SUMMARY
II. INTRODUCTION
A. The Draft HARS Framework Report
B. Remediation Material Workgroup Charge to Peer-Reviewers
C. Scope of Technical Comments
D. The Difficulties Evaluating Sediment Conditions
III. GENERAL COMMENTS
A. Clarity and Transparency of the Proposed HARS Framework
B. Current Environmental Conditions at the HARS
C. Comparisons to a Reference Area
D. "Current HARS-Specific Guidelines"
IV. RESPONSES TO SPECIFIC RMW CHARGES
A. Overall Process
B. Bioaccumulation Testing Analytes
Bl. Proposed Additions to Analyte List: Alkylated PAHs
B2. Proposed Additions to Analyte List: Organotins
B3. Proposed Additions to Analyte List: Coplanar PCB Congeners
C. Comparison to Reference
D. Adjustment to Steady State
Dl. Adjustment to Steady State: Organic compounds
D2. Adjustment to Steady State: Metals
E. Human Health Evaluations
El. Overall
E2. Human Health Evaluations: Comparison to HARS-Specific Values
E3. Human Health Evaluations: Consideration of Combined Effects
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TABLE OF CONTENTS (Cont'd)
V. REFERENCES CITED
LIST OF FIGURES
Figure 1. USEPA Region 2 / USACE-NYD interim (current) HARS framework for
evaluating bioaccumulation test results for project sediments.
Figure 2. Proposed USEPA Region 2 framework for evaluating human health risks
associated with bioaccumulation test results.
Figure 3. Questions and potential shortcomings with USEPA Region 2's proposed
framework for evaluating bioaccumulation test results for project sediments.
Figure 4. Overview of the complete proposed USEPA Region 2 framework for
evaluating the suitability of sediments for use as remediation material at the
HARS.
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I. EXECUTIVE SUMMARY
Under the direction of Battelle, a Scientific Peer Review Panel was charged on January 8-
9, 2002 with review of the December 21, 2001 draft report prepared by the United States
Environmental Protection Agency (USEPA) Region 2 titled, "Proposed Bioaccumulation
Testing Evaluation Framework for Assessing the Suitability of Dredged Material to be
Placed at the Historic Area Remediation Site (HARSV' (hereafter referred to as the
HARS Framework report). The Scientific Peer Review Panel was asked to provide
technical comments on thirty questions and concerns developed by USEPA Region 2 and
the Remediation Materials Workgroup (RMW) pertaining to evaluation of the potential
human health risks associated with evaluation of the suitability of dredged sediment for
use as remediation material at the HARS. Questions and concerns pertaining to
ecological issues will be addressed in a second phase of the peer-review process.
Technical Findings
Overall, USEPA Region 2 is to be congratulated for undertaking a process to revise the
current interim program for evaluating the suitability of dredged material for placement at
the HARS. The Agency has developed a sound foundation for further refinement of a
dredged material evaluation framework for the New York / New Jersey (NY/NJ) Harbor
region. It is evident that the Agency strove to develop a scientifically defensible
evaluation process. The Agency succeeded in several aspects, but fell short, most
notably, with regard to the calculation of HARS-Specific values and human health risks.
Additional refinements of the proposed HARS Framework are needed to ensure that the
most current understanding of fate and effects processes and the most advanced practical
procedures for evaluating contaminant migration through the food web and human health
risks are included in the final framework.
The proposed HARS Framework raises four general concerns, which USEPA Region 2 is
urged to consider and address in the final dredged material evaluation framework. These
concerns include the need for greater clarity and transparency in the evaluation test
procedures and pass/fail criteria. A critical deficiency at the forefront of this concern is
the absence of any description of the approach for conducting a weight-of-evidence
review. The Agency does not specify a scheme for decision-making when the results of
one or more Green Book sediment toxicity tests, statistical comparisons, and risk
evaluations fall in the gray area between clear pass and fail criteria.
A second general concern is the characterization of current environmental conditions at
the HARS based on insufficient and/or potentially outdated evidence concerning the
status of the fishery, recreational angler activity and catch success, contaminant residues
in benthic organisms and fish tissues, geochemistry, and the fate and effects of
contaminants placed at the HARS. Given the considerable resources invested in future
remediation efforts, as well as the opportunity to potentially improve conditions at the
HARS while at the same time supporting port/harbor development and the region's
economy, it is not unreasonable for the Agency to implement a HARS-specific sediment
and biological characterization study and monitoring program to establish current,
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baseline environmental conditions and to continually monitor for future changes. At
present, there appears to be little credible evidence upon which to ascertain whether (and
how) the placement of dredged sediments will improve conditions (e.g., reduce
contaminant body burdens in fish) at the HARS.
The third and fourth general concerns address whether the reference area selected for
comparing bioaccumulation test results in project sediments is appropriate, and the
significant uncertainties associated with evaluating the impact of contaminants in
sediment on human health and wildlife. The physical characteristics of reference
sediments appear to contrast sharply with sediment conditions at the HARS and the
physical characteristics of the majority of sediments in the NY/NJ Harbor region. Given
the importance of sediment physical and chemical properties on contaminant
bioavailability, this seemingly biased evaluation benchmark only compounds the already
significant uncertainties associated with integrating sediment chemistry analysis, toxicity
testing, bioaccumulation tests, tissue analysis, and risk assessment into sediment
evaluation framework.
With regard to responses to the thirty questions and concerns raised by the RMW
charges, several important themes emerged in the course of this review. First, the HARS
Framework could be greatly improved by using HARS-specific information in the risk
evaluation framework. At present, the proposed HARS Framework appears to
overemphasize the maximally exposed individual through repeated use of several
conservative exposure factors and contaminant fate and effects assumptions, which may
dictate a negative evaluation outcome for most, if not all, project sediments.
Second, adoption of a probabilistic approach to several aspects of the evaluation
framework could help resolve the five areas of uncertainty that appear to be the focus of
considerable concern among members of the RMW. These uncertainties appear to
include: (a) contaminant levels in project sediment; (b) current environmental conditions
at the HARS; (c) relevance of test organisms in bioaccumulation tests to organisms that
inhabit the HARS; (d) food web transfers of contaminants in sediment to prey and to
upper trophic level fish caught and consumed by recreational anglers; and, (e) fishing and
consumption habits among different angler populations.
How can the Agency overcome or manage these uncertainties? A probabilistic approach,
which is typically undertaken using a Monte Carlo statistical method, addresses the
deficiencies evident throughout the proposed HARS Framework because of its reliance
on a point estimate (i.e., deterministic) approach. A probabilistic approach is capable of
incorporating the range of known or suspected values for key environmental and
exposure factors, and can impart a great deal more information on the results of the risk
evaluation to the Agency and stakeholders. The Agency is strongly encouraged to
consult guidance documents available from several USEPA sources and to adopt a
probabilistic approach as a first step to addressing many of the technical deficiencies
identified in this review.
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Going forward, the Agency should further refine the suite of sediment assessment tools
indicated in the HARS Framework to evaluate chemistry conditions and biological
impacts associated with project sediments. The Agency also should clearly define a
weight-of-evidence scheme for evaluating these data. Every effort in the future should
focus on the introduction of HARS-specific environmental and exposure factors into the
data compilation, risk assessment and sediment evaluation processes. The results of these
activities will greatly improve the proposed HARS Framework. The Agency should
recognize that whatever final form the framework takes, the process will likely need to be
revised again in the future, as more information becomes available. It is expected that
advances in the scientific understanding of chemical fate and effects in the marine
environment will continue. USEPA Region 2 should be prepared to assess and adopt
new information and evaluation methods, as appropriate, in the future.
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II. INTRODUCTION
Under the direction of Battelle, a Scientific Peer Review Panel (Panel) was charged with
performing a technical review of the December 21, 2001 draft report prepared by the
United States Environmental Protection Agency (USEPA) Region 2 titled, "Proposed
Bioaccumulation Testing Evaluation Framework for Assessing the Suitability of Dredged
Material to be Placed at the Historic Area Remediation Site (HARS)." In consultation
with the Remediation Materials Workgroup (RMW), USEPA Region 2 developed thirty
questions directed at resolving specific areas of concern or controversy pertaining to the
proposed framework for evaluation of potential human health risks associated with
contaminants bioaccumulated from dredged material proposed for use as remediation
material at the HARS. This is the first of a two-phased peer-review process; the second
phase will address the ecological implications of the proposed HARS framework.
A. The Draft HARS Framework Report
The history of dredged material disposal in the New York/New Jersey (NY/NJ) Harbor
Region is described in USEPA (1997). Briefly, the New York Bight Apex has been used
for disposal of dredged material, municipal garbage, building debris, sewage sludge, and
other waste materials since the 1880's. The New York Bight Apex is defined as an
approximately 2,000 km2 area within the larger New York Bight extending along the
northern coast of New Jersey from Sandy Hook southward to 40° 10' latitude and
eastward along the Long Island from Rockaway Point to 73° 30' longitude.
During the past nearly 25 years, ocean disposal of sediment from NY/NJ Harbor has been
restricted within the New York Bight Apex to an area referred to historically as the Mud
Dump Site (MDS). The MDS was designated as an interim site for dredged material
disposal following enactment of the Marine Research, Protection and Sanctuaries Act
(MPRSA) in 1972. The MDS received final site designation in 1984 for disposal of up to
100 million yds3 of dredged material from navigational dredging projects associated with
the Port of NY/NJ and nearby harbors.
In February 1995, USEPA Region 2 initiated an environmental review process with
regard to the future of the MDS, and in May 1997 proposed to de-designate and terminate
ocean disposal of dredged material at the MDS. Simultaneous with closure of the MDS,
the site and surrounding area historically used as for disposal of dredged materials would
be redesignated under 40 CFR 228.11 of the MPRSA as the HARS. The intent of the
redesignation was to manage the area formerly known as the MDS to reduce the impacts
of historical disposal activities to acceptable levels.
In accordance with 40 CFR Section 228.11(c), the remediation of the HARS would
involve the use of uncontaminated dredged material, which was defined as dredged
material that meets Category I standards and would not cause significant undesirable
effects including through bioaccumulation. Dredged material meeting these criteria was
referred to as "the material for remediation" or "remediation material". Under previous
guidance within USEPA Region 2 and the United States Army Corps of Engineers-New
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York District (USACE-NYD), sediments proposed for placement at the MDS and
classified as Category I material represented material allowed for ocean disposal without
capping. Category II material required capping and Category III material was prohibited
from ocean disposal. Under the terms of the HARS designation, only Category I material
would be suitable for use as remediation material. Among the technical issues of concern
to both USEPA Region 2 and the USACE-NYD with regard to the HARS is the
development and implementation of a framework for determining the suitability of
dredged sediment as Category I material.
B. Remediation Materials Workgroup Charge to Peer-Reviewers
The RMW reviewed the process that USEPA Region 2 is considering for evaluation of
potential human health risks associated with contaminants bioaccumulated from dredged
material proposed for use as remediation material at the HARS. Members of the RMW
include representatives from USEPA Region 2, USACE-NYD, the states of New York
and New Jersey, and other New York Harbor port and environmental groups.
As a result of RMW discussions, a set of thirty questions and concerns were developed
for consideration by a Scientific Peer-Review Panel to resolve specific areas of concern
or controversy regarding the sediment evaluation process. The following questions and
concerns were posed to the Panel:
Overall Process
1. Throughout the proposed process, there are various uncertainties introduced.
Please identify the key areas of uncertainty that need to be addressed. Are there
additional data sources or parameters that could be used to address these areas?
What methods are available for describing and accounting for these uncertainties
in the calculation of HARS-Specific Values? Of the methods available, which
would you recommend for consideration and why? Please consider the
implications of implementing these methods in the regulatory framework. Please
include an evaluation of probabilistic and deterministic methods in your
discussion.
Bioaccumulation Testing Analvtes
2. Is measurement of the 16 priority pollutant PAHs (i.e., parent PAHs) sufficient
for characterizing the risks associated with the total PAH bioaccumulated by
organisms exposed to dredged material proposed for placement at the HARS?
Does measurement of the alkylated compounds significantly improve risk
assessment of PAHs?
3. Is the proposed adaptation of EPA Method 8270 (Appendix D) acceptable and
appropriate for regulatory decision-making? If not, what is an acceptable and
appropriate method?
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4. Under what specific conditions would the testing for alkylated PAHs for a
particular project be appropriate and warranted?
5. What uncertainties would be introduced within the analysis of risk should
alkylated PAHs be included? What steps could be taken to account for these
uncertainties in decision-making? Given the likelihood the method for using non-
detects (as described in EPA/CENAN, 1997) will result in an overestimate of risk,
what are the implications?
6. It is recognized that additional methods have been used for the analysis of
organotins (e.g., Krone et al., 1989). Will the proposed analytical method (Rice et
al., 1987) provide adequate data of sufficient quality to assess relevant risks from
organotins? If not, please provide recommendations.
7. What special QA/QC procedures should be implemented to ensure the quality and
usability of the organotin data?
8. Under what specific conditions would the testing for organotins for a particular
project be appropriate and warranted?
9. If the approach for evaluating dioxin is modified, should it include the
contribution of PCBs with dioxin-like activity as proposed? If so, how?
Comparison to Reference
10. Please consider the policy for assigning values (at one half the detection limit) to
tissue residues that are reported as "
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14. For the DDT derivatives and dieldrin, please comment on the appropriateness of
using M. nasuta data rather than N. wVeMs-specific data in the estimation of steady
state multipliers.
15. Are the approaches taken to adjust organic contaminant bioaccumulation data to
steady state adequate? Do the proposed multipliers agree with previously
published studies (i.e., do they appear reasonable)? If not, please elaborate.
16. What are the major sources of uncertainty associated with the approaches? What
alternative approaches would reduce the uncertainties? How could these
uncertainties be described and accounted for in decision-making?
17. In your opinion, is the methodology followed to derive the steady state multiplier
for non-essential metals (i.e., a factor of three) scientifically appropriate
(Appendix G)? Please elaborate. Do you have any recommendations of additional
or alternate methodologies or information that can be used to either supplement or
replace the proposed method?
Human Health Evaluations
18. Please comment on each factor listed above (and in Table 5) as to its
appropriateness for use in the equations listed above. Would you recommend
additional factors? Would you change or modify the equations as written above?
If so, how?
19. Are the methods used to derive the human health exposure parameters and
assigned values discussed in Section E appropriate (please review the referenced
appendices)? If not, please elaborate. How should these factors be factored into
the risk analyses and decision-making?
20. Is the approach taken to relate fish whole body and fillet concentrations
scientifically appropriate? If not, what method would you recommend?
21. Could the analysis be improved by focusing on key fish (seafood) species at the
HARS? What characteristics should be used to select these key species?
22. In your opinion, is the approach for assuming total metal to be in the most toxic
form appropriate and reasonable? Should metal speciation/complexation be
considered in the assessment of metals bioaccumulation, trophic transfer, and
human health risks? Is the proposed approach for evaluating methyl mercury
appropriate? Are there alternative analytical or risk assessment techniques
available that would improve the risk assessment of metals? Is the multiplier
proposed for adjusting measured concentrations of arsenic appropriate and
reasonable?
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23. Is the assumption that the potency of alkylated PAHs can be estimated by the
potency of the parent PAH appropriate? Is this assumption likely to result in an
under- or overestimate of the risk associated with the alkylated PAHs?
24. Please comment on the potential for human exposure to PAHs through
consumption of finfish and other seafood.
25. What are the major sources of uncertainty associated with the approaches
described in Section E? What alternative approaches would reduce the
uncertainties? How could these uncertainties be described and accounted for in
decision-making?
26. What is your recommendation for evaluating the potential toxicity of organotins?
Should they be evaluated as individual compounds? Summed as total? Should
there be some consideration of relative toxicity?
27. Please comment on the appropriateness of the proposed approach for converting
and using the analytical data for alkylated and parent PAHs to estimate risk from
all PAHs.
28. Do you believe that the "disaggregate" modeling discussed above (and shown in
Figure 4) for estimating human health HARS-Specific Values for lead is
appropriate? Would you recommend an alternative risk assessment method be
used given the information and data available? Do you believe the method
described has appropriately taken uncertainty into account? Please elaborate.
29. In your opinion, are the methodologies and equations described above appropriate
for estimating total carcinogenicity and combined non-cancer impacts of
contaminant mixtures accumulated from dredged materials proposed for use as
Remediation Material at the HARS?
30. Is the conceptual model for evaluating fish exposure to dredged material at the
HARS and human exposure through ingestion of seafood appropriate and
reasonable? How can the uncertainties associated with the assumptions in this
conceptual model be reduced? Please consider the spatial and temporal elements
of exposure in your discussion.
C. Scope of Technical Comments
As a member of the Scientific Peer-Review Panel, my responses to the thirty questions
and concerns raised by the RMW are contained in this document. Section III summarizes
my general overall comments and technical concerns on the proposed HARS Framework
process that, in my opinion, warrant further attention by USEPA Region 2. My responses
to the thirty specific questions and concerns identified by the RMW are presented in
Section IV.
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DRAFT
in. GENERAL COMMENTS
USEPA Region 2 is to be congratulated for undertaking a process to revise the current
interim program for evaluating the suitability of dredged sediments for use as remediation
material for the HARS. This initiative is much needed, in light of the advances in the
state of the sciences relating to understanding the fate of chemicals in aquatic
environments and their effects on human health and wildlife, as well as advances in
computer models of these processes and statistical methods for performing risk
assessments.
Overall, the Agency has developed a sound foundation for further refinement of a
dredged material evaluation framework. It is evident that the Agency strove to develop a
scientifically defensible evaluation process. The Agency succeeded in several aspects,
but fell short, most notably, with regard to the calculation of HARS-Specific values and
human health risks (i.e., Section E of the proposed HARS Framework). Additional
refinements of the proposed HARS Framework are needed to ensure that the most current
understanding of fates and effects processes and the most advanced practical procedures
for evaluating contaminant migration through the food web and human health risks are
included in the final framework.
However, it is likely that whatever final form the HARS Framework takes, the evaluation
process will need to be revised again in the future, as more information becomes
available. It is expected that advances in our understanding of chemical fate and effects
in the marine environment, the status of various fisheries in the New York Bight Apex,
and recreational angler and commercial fishing activity at the HARS will continue.
USEPA Region 2 should be prepared to assess and adopt these advancements and other
new information, as appropriate, in the future.
My general comments focus on the following concerns:
A. Clarity and Transparency of the Proposed HARS Framework
Presumably, USEPA Region 2 will prepare a complete and thorough guidance document
that encompasses all of the requirements that an applicant must take into consideration
when proposing dredged material for placement at the HARS. Owing to the complexity
of the various technical subjects pertinent to the proposed HARS Framework, it is
incumbent on the Agency to prepare a clear and comprehensive package of information
using the typical USEPA guidance format.
As currently written (and, perhaps only for the purposes of this peer-review), the
document was very difficult to follow and the information inconsistent with the overall
approach presented in Figure 1 (pg.10 of the peer-review document). A source of
additional confusion was the difference between the proposed evaluation framework
depicted in Figure 1 (pg. 10 of the peer-review document), the text on pages 7-8, and the
current, interim framework shown in Appendix B. Both frameworks are reproduced in
Figures 1 and 2. According to pgs 7-8 and either missing or incomplete in Figure 1 (pg.
12
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RJ. Wenning Scientific Peer Review - HARS Evaluation Framework
Chemical "X"
Greater than
Reference?
b,
I
^No
Do Matrix levels
or Regional
Dioxin value
for Chemical "X"
Exist?
Yes
Chemical "X"
greater than
Matrix level?
Dioxin value
greater than
category u level.'
No
Individual Chemical
Effects Evaluation
No
Yes
Chemical "X"
Yes
greater than
~
FDA levels?
Dredged Material
exceeds LPC and
is not suitable for
ocean placement
No
Risk Evaluation
for Chemical
"X?"
Dredged
Material is not
Categoiy 1.
Yes
J
Integrated Effects
Evaluation of
Bioaccumulation
Results Using 8
Green Book
Factors
2Remediation Material
determination
'Letter refers to discussion section in
document
^Note: If any chemical exceeds an FDA
Action Level, a Matrix level or Dioxin
Category lvalue, dredged material is
not Category 1.
Note: A revised risk-based Matrix value
of 113 ppb for PCB in worm has been
adopted.
Figure 1. USEPA Region 2/USACE-NYD interim (current) HARS Framework for
evaluating bioaccumulation test results for project sediments, (from: USEPA Scientific Peer-
Review Package and Charge, December 21,2001, Appendix B).
13
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RJ. Wemiing Scientific Peer Review - HARS Evaluation Framework
Chemical "X"
measurement
U)Is
chemical "X"
gnater ih in
reference?
Yes,
r
(2a i Is
chemical *'X"
dioun'
Yes
r
(2b) Does
t heiniial
"X" exceed
dm tin
No
Yes
Individual Chemical
Effect Evaluation
No
No
¦\d]USt tO
steady state,
if ncLCs-jr}
\djuit to
steady st ite
(3) In comparing chemical \ to its
HARS-Specilic \ dlue, lIols flic
weight-ot-e\ ide iit_e determine
there to be a potential tor
significant undesirable cHllIs
through its bioaicumuLiuon in
accoidamc \Mth 40 ( 1 R 227 <»'
Potentially
Suitable
For use as
Remediation
Material
Not
Remediation
No
Yes
Remediation
Material
No
(4) I)i es the weight of evidence
regarding the ('ombincd
Effects Evaluation (i.e,
utmpaiison ot dredgci matei uil
as a whole to (1JK, loUl
lOKiiiiigeniul. and noil ljiillt
hazard index) letermme there
lo hi. a potential for
sigiiitkant undesirable elteus
through hioaciumulaUon in
aimrdaiKe w ilh 40 CI R 227 6'
Yes
Not
Remediation
Material
Figure 2. Proposed USEPA Region 2 framework for evaluating human health risks
associated with bioaccumulation test results, (from: USEPA Scientific Peer-Review Package
and Charge, December 21,2001; Proposed TEF, p.10).
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R.J. Wenning Scientific Peer Review - HARS Evaluation Framework
10 of the peer review document), the current HARS-specific guidelines involves the
following considerations:
1. Residues of 5 pesticides, PCBs, and mercury in sand worm (Nereis virens) and
bent-nosed clams (Macoma nasuta) tissues exposed to sediment in 28-day solid-
phase bioaccumulation tests must be below FDA Action Levels.
2. Regional matrix values for mercury, cadmium, total PCBs, and total DDT
established in 1981 are not exceeded in sand worm (Nereis virens) and bent-nosed
clams (Macoma nasuta) tissues exposed to sediment in 28-day solid-phase
bioaccumulation tests.
3. The wet weight tissue concentration of 2,3,7,8-TCDD does not exceed 1 part per
trillion (pptr) and the toxic equivalence of all non-2,3,7,8-TCDD dioxin and furan
congeners does not exceed 4.5 pptr in sand worm (Nereis virens) and bent-nosed
clams {Macoma nasuta) tissues exposed to sediment in 28-day solid-phase
bioaccumulation tests.
4. Contaminants in sediment not having regional matrix values or dioxin values do
not exceed human health guidelines [or ecological benthic tissue guideline values]
using standard USEPA default exposure assumptions (e.g., 70 kg body weight,
6.5 g of fish consumed per day over a 70-year lifetime) at 10"4 cancer risk and
hazard quotient less than 1 for non-cancer effects.
The proposed framework diagram should specify several steps that the Agency appears to
retain from the current, interim evaluation approach such as comparison to Regional
Matrix Values and FDA Action Levels. The consideration of chemicals other than dioxin
is missing, and dioxins appear to receive an inordinate level of attention relative to the
breadth of HARS analytes. The actual risk assessment considerations are confined to a
small portion of the framework, and may rarely come into play given the large handicap
that must be overcome at the outset when bioaccumulation test results for project
sediments are compared with those for reference sediments. These questions and other
potential shortcomings with the proposed framework are highlighted in Figure 3.
Figure 4 describes this reviewers understanding of the complete proposed HARS
Framework, including USACE-NYD Green Book sediment toxicity testing requirements.
The proposed HARS Framework also includes the following additional requirements: (a)
acute toxicity test results involving 3 sensitive marine organisms exposed to the liquid
phase of project sediments should not exceed a toxicity threshold of 0.01 of the
concentration shown to be acutely toxic; (b) contaminant concentrations in the liquid
phase of project sediments should not exceed marine water quality criteria; (c) acute
toxicity test results involving inland silversides (Menidia beryllina), mysid shrimp
(Mysidopsis bahia) and blue mussel (Mytilus edulis) exposed to the suspended particulate
phase of project sediments should not exceed a toxicity threshold of 0.01 of the
concentration shown to be acutely toxic; (d) acute toxicity test results involving mysids
and amphipods exposed for 10-days to the solid phase of project sediments should not
exceed mortality in reference sediments by 10% for mysids and 20% for amphipods and
15
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R. J. Wenning Scientific Peer Review - HARS Evaluation Framework
jirncat
dhUlCt
chemical "V
1 grcator than
| reference?
na)h
p chemical
I diuxin.' ~
~
Hipii®
Chemical
iftllSiS#
cfioxtn
Project sediments are
practically doomed to
fail when compared to
reference sediments!
lea!
mm rt
No
to
Dioxin appears to plav
an inordinatel> large
role in the evaluation!
v - ;
steady state
fitltllll
Here is the risk . '
assessment! But, ,
where (and how) does
weight-of-evidence
enter the decision-
making process?
Afjghl '>t C\ idui-C !•!>!!
c'lgiiificrint undebifaWc eft'-vta
shiough its hioiicct'mulaiimt in
IplpgJteillpilll.
icgaidiib: the Combined
comparison ot'druked material^
as a v-holc 10 (BR. total
carciiiORcnscity i>fid non-caiiccr
lu/ard index) determine tberc ~ -"
What about other
chemicals? Regional
matrix \ alues? FDA
Action Levels?
Not
iienk'diai.ioi;:.
Figure 3. Questions and potential shortcomings with USEPA Region 2's proposed HARS
Framework for evaluating bioaccumulation test results.
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DRAFT
Green Book Framework
Fail
HARS Framework
Pass
Fail
Fail
Pass
Pass
Fail
Pass
Fail
Pass
Fail
Pass
Solid phase testing
Acute tox tests, 3 species
Acute tox tests, 3 spcci
Acute tox tests, 2 species
Compare to marine WQC
Liquid phase testing
Suspended particulate phase testing
Bioaccumulation tests, worm & clam
Compare to Reference sediment test
Compare to HARS-Spccific Values
Compare to FDA Action Levels.
Compare to Dioxin Matrix Values
Compare to Regional Matrix Values
Suitable
for use at
HARS
Weight of
Evidence
review
Not suitable
for use at
HARS
Compare to fish consumption
human health risk assessment
(fish consumption only)
Figure 4. Overview of the complete proposed USEPA Region 2 framework for evaluating the
suitability of dredged sediments for use as remediation material at the HARS.
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should not be statistically greater than reference sediment results; (e) correction of 28-day
bioaccumulation tests for steady-state where results indicate 28 day exposures may not be
sufficient to represent field conditions after placement; and (f) comparison of tissue levels
to background tissue concentrations in the general area of the HARS.
A critical deficiency at the forefront of the issue of clarity and transparency is the absence
of any description in the proposed HARS Framework of the approach for conducting a
weight-of-evidence review. Although indicated in the peer-review document (Figure 2),
the Agency does not describe, in any detail, a scheme for decision-making when the
results of one or more Green Book sediment toxicity tests, statistical comparisons, and
risk evaluations fall in the gray area between clear pass and fail criteria. Figure 4 was an
attempt to capture the overall HARS Framework proposed by USEPA Region 2 and to
identify where a weight-of-evidence step might occur in the proposed evaluation process.
The Agency is strongly encouraged to revise this diagram to specify where (and how) a
weight-of-evidence evaluation would be conducted.
In summary, the evaluation framework sorely requires presentation in a logical step-wise
fashion. The hypothetical example of the current evaluation framework presented in
Appendix B was very helpful, and contributed greatly to understanding the approach the
Agency was striving to revise or improve. A clear, and accurate big picture of the
proposed evaluation framework is needed. The clarity and transparency of the
framework approach and the technical information used to support each evaluation step
must be improved.
B. Current Environmental Conditions at the HARS
The HARS Framework is based on several assumptions regarding current environmental
conditions at the HARS for which insufficient and/or potentially outdated evidence is
presented in Appendix A and elsewhere in the peer review document. According to the
HARS supplemental environmental impact statement (SEIS) and USEPA's Response to
Comments on the May 13,1997 proposed rule, the decision to remediate the HARS using
Category I sediment as capping material was based on: (a) the presence of toxic effects in
the HARS; (b) tissue residue data indicating dioxin bioaccumulation exceeding Category
I levels in worm tissue collected from the HARS; and, (c) PCB/TCDD contamination in
area lobster stocks.
This reviewer does not have an opinion with regard to the conclusion that the HARS
should be remediated. However, the Agency's position (which is implicit in the peer
review document and explicit in Appendix A) that current environmental conditions at
the HARS are understood is flawed for several reasons. Specifically:
• The results of 10-day amphipod (Amplesica abdita) bioassay tests of sediments from
the HARS are at least 8 years old. Bioassay data from 1994 do not provide an
indication of current sediment conditions today, or to what degree contaminants in
sediment at the HARS bioaccumulate and are responsible for the levels reported in
fish and shellfish. Using this information to represent baseline conditions for the
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purposes of monitoring future environmental improvements at the HARS is
inappropriate.
• Tissue analyses for contaminant residues by Pruell et al. (1990) of benthic biota
(twelve worms and five clams) collected in 1987 and 1989 from eight locations, four
of which may lie outside the HARS, provides little meaningful information either on
current body burdens in benthic organisms indigenous to the HARS or food chain
transfers. Using this information to represent baseline conditions for the purposes of
monitoring future environmental improvements at the HARS is inappropriate.
• Evaluation of contaminant concentrations in HARS sediments using Long et al.
(1995) effects range low and median (ER-L, ER-M) provide little meaningful in
information, even for screening purposes. Nearly all of the chemical-specific ER-L
and ER-M values are derived from datasets that combine test results from freshwater
and marine environments and different species, several of which may not be present
in the HARS. While the approach allows for interpretation of possible effects across
a range of concentrations, it does not address the fundamental issue of establishing
the link between occurrence in sediment and adverse effect.
• Contaminant conditions in lobster stock from the New York Bight provide little
useful information pertaining to current conditions in the population that may inhabit
or forage within the HARS. Results from studies of lobsters from smaller
geographical areas such as coastal harbors in Maine indicate that tissue residues
typically reflect broader, rather than site-specific, environmental conditions (Wenning
et al, 1998). The source of contaminants is nearly impossible to discern in lobster.
The Agency correctly notes that the migratory behavior of lobsters confounds
definitively linking specific areas of dredged material disposal, to other dumping
activities, or to other ongoing pollution sources (Appendix A, pg. 35).
Given the considerable resources invested in future remediation efforts, as well as the
opportunity to potentially improve conditions at the HARS while at the same time
supporting port/harbor development and the region's economy, it is not unreasonable for
the Agency to implement a HARS-specific sediment and biological characterization
study and monitoring program to establish current baseline environmental conditions and
to continually monitor for future changes. This would include other necessary or
appropriate considerations specified in 40 CFR 228.10(b) important to understanding to
what extent the marine environment has been impacted by historical activities, as well as
future disposal activities. These additional considerations include (a) movement of
materials into estuaries and sanctuaries, oceanfront beaches and shorelines; (b) movement
of materials toward productive fishing grounds; (c) absence of pollution -sensitive
species characteristic of the general area; (d) non-seasonal changes in water quality or
sediment composition; and, (e) non-seasonal changes in pelagic, demersal, or benthic
biota in the HARS.
In summary, there appears to be little credible scientific evidence to understand current
sediment and biological conditions at the HARS. Furthermore, there appears to be a
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weak foundation for understanding whether (and how) the placement of dredged
sediments will improve (e.g., reduce contaminant body burdens in fish, reduce exposure
to recreational anglers, improve the quality of the region's fishery) at the HARS. As
such, the Agency does not appear to be in a position to determine whether remediation
material designated for the HARS will support the stated risk management goal of
improving current conditions at the HARS. The Agency is strongly encouraged to begin
development of a coordinated, scientifically defensible investigation of sediment and
biological conditions at the HARS prior to implementation of the evaluation process
described in the proposed HARS Framework. The significance of this information is
critical to determining the future success of restoration activities at the HARS.
C. Comparisons to a Reference Area
The reference area identified in the proposed HARS Framework for comparing
bioaccumulation test results in project sediments and, presumably, to represent future
desired conditions at the HARS is inappropriate. The proposed HARS Framework
indicates that reference sediment is predominately (95%) sand collected from an area
located south of the HARS. The physical characteristics of reference sediments appear to
contrast sharply with sediment conditions at the HARS and the physical characteristics of
the majority of project sediments in the NY/NJ Harbor region (see e.g., Huntley et al.,
1994), which are likely to contain predominately fine grained, organic rich silts and clays.
It is widely reported in the scientific literature that fine grained materials rich in organic
material and other constituents such as organic ligands and inorganic oxides and sulfides
control the bioavailability of contaminants (Word and Word, 2001; USEPA, 2001). The
toxicity and bioaccumulation potential of contaminants is generally considered to be
significantly lower in silty harbor and navigation channel sediments as compared to
coarse-grained, sandy organic-deficient materials such as encountered in the reference
area.
Currently proposed as one of the first evaluation steps in the proposed HARS
Framework, it would appear to be an almost insurmountable criterion for most project
sediments to achieve for consideration as remediation material. Consequently, the
current reference area selected by USEPA Region 2 is inappropriate. The reference area
is inconsistent with current USEPA guidance for selecting a reference area for risk
assessment and remediation purposes (USEPA, 1997). Because the identification of the
Reference Site is critical to understanding baseline conditions at the HARS and for
establishing benchmarks for defining successful remediation, the Agency should
reconsider its Reference Site.
Given that the HARS encompasses a 30 square mile area, has the Agency considered
developing a sediment database comprised of data from sediments collected within New
York Bight Apex? Data on different environmental factors (e.g., sediment grain size,
organic carbon content, ammonia levels, etc) could be used to calculate a distribution of
physical characteristics from which to identify a representative "reference" or
"background" condition.
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D. The Difficulties Evaluating Sediment Conditions
Despite the Agency's positive efforts, there continue to be significant uncertainties on
how to best integrate the data generated from sediment chemistry analysis, toxicity
testing, bioaccumulation studies, fish/wildlife tissue analysis, and risk assessment to
evaluate sediment conditions and determine the risks to human health and wildlife
(Wenning, 2001; AFS, 2000; Ingersoll et al, 1997; Solomon et al, 1997; Engler, 1980;
Chapman, 1989). Several potentially confounding factors affect the interpretation of both
field measurements and laboratory test data, which raises additional uncertainties when
extrapolating or predicting environmental fate and toxic effects to either humans or
wildlife (DelVails et al., 2002; Word and Word, 2001).
The key concerns raised among scientists and environmental regulators include, but are
not limited to: the relative degree of agreement (concordance) among different
assessment methods, the ability to demonstrate the absence, or presence, of toxicity of
chemicals in sediment to relevant target organisms; the capability to identify cause and
effect relationships; the applicability of individual assessment methods to different
sediment types or environmental conditions; and, the ability to minimize uncertainties
and limitations (Chapman and Wang, 2001; USEPA, 2001; O'Connor and Paul, 2000;
Ankley, 1997).
Among these concerns, the different sediment assessment benchmarks used to evaluate
sediment chemistry, tissue residue, and sediment toxicity data provide no clear cause and
effect relationships upon which to conclude whether the occurrence of contaminants
poses a threat to humans (Dickson et al., 1987). Each of the benchmarks is conservative,
and lack the necessary and relevant site-specific environmental and human exposure
information upon which regulatory sediment management decision-making should rely.
As stated earlier, before adopting the proposed HARS evaluation process, a clearly
defined weight-of-evidence methodology supported by the scientific community is
strongly recommended.
Going forward, the Agency should continue to emphasize the assimilation of relevant
chemistry and biological data, further refine the suite of sediment assessment tools to
evaluate these data, and the weight-of-evidence scheme for classification of sediment
conditions. Every effort in the future should focus on the introduction of site-specific
factors into the data compilation and evaluation processes. The results of these activities
will greatly improve the opportunity to provide a technically sound basis for future
assessment and remediation activities at the HARS.
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IV. RESPONSE TO SPECIFIC RMW CHARGES
As a member of the Scientific Peer Review Panel charged with review of the December
21, 2001 draft proposed HARS Framework, below are my answers to the thirty questions
and concerns raised by the RMW pertaining to evaluation of the potential human health
risks associated with evaluation of the suitability of dredged sediment as remediation
material at the HARS.
A. Overall Process
J Ihroughoul the proposedproiess, rhirt are li/mrtf unierhunm s inttodund Phase identify
the key areas oj uiiLcriaintv thui need to be addressed .be there additional dura source* or
parameters that could he used to address tlwse aieas.' What methods are available far
describing and atc-runiing for tht unctHaintu s m tht t alt ulatlun of IIItf.vyt i iflc Vahu \1
Of Uil m
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R.J. Wenning Scientific Peer Review - HARS Evaluation Framework
The proposed HARS Framework includes 9 metals, 15 chlorinated pesticides, 16 parent
polycyclic aromatic hydrocarbons (PAHs), 22 polychlorinated biphenyls (PCBs)
congeners, and 17 2,3,7,8-substituted dioxins and furans in bioaccumulation testing
requirements for project sediments. The list of chemicals was developed based on a
review of available data from Squibb et al. (1991). Total PCBs in tissue residues are
estimated based on two times the sum of the measured concentrations of 22 individual
PCB congeners. USEPA Region 2 is proposing to add alkylated PAHs and organotins to
the list of analytes in the HARS Framework.
2 l\ mt.asurt.ment o1 the lf> priority pollutant PUl\ (i e patent HAlls) \uffitwiu
characterizir.g tha rusks associated with the total HAH hioaccumulated by organisms
to dredged material proposed for placement at the HARS? Does measurement of the alkyh
PAHs are ubiquitous in the environment (Eisler, 2000). Of major environmental concern
are the mobile PAHS that vary widely in molecular weight from CIO to C24 compounds.
Higher molecular weight PAHs are relatively immobile, with low volatility and relatively
in soluble in water. The lower molecular weight PAHs containing two to three rings such
as the naphthalenes, fluorenes, phenanthrenes and anthracenes are acutely toxic to some
organisms, whereas the higher molecular weight (four to seven ring aromatic compounds)
generally are not acutely toxic. However, all of the known or suspected carcinogens, co-
carcinogens, and tumor promoters are in the high molecular weight PAH group (Eisler,
2000; ATSDR, 1999; USEPA, 1999). Both low- and high- molecular weight PAHs are
known to occur in Newark Bay, New Jersey and elsewhere in the NY/NJ Harbor region
(Huntley et al. 1993,1995).
Including alkylated PAHs in the list of HARS analytes appears to be warranted to a
limited degree for both ecological and human health risk assessments. PAH
accumulation has been extensively investigated in molluscs and fish, and, to some extent,
in lobsters (Eisler, 2000; USEPA, 1980). It appears that the more water-soluble, lower
molecular weight PAHs bioconcentrate to a higher extent than the higher molecular
weight PAHs, although significant and rapid depuration is known to occur when animals
are removed to a clean environment (Eisler, 2000; Baumard et al., 1998). Similar
observations are observed in lobster. PAH levels in fish are usually low due to rapid
metabolism and limited bioaccumulation (Eisler, 2000; USEPA, 1980).
From strictly a human health risk assessment standpoint, characterization of the risks
associated with parent PAHs should be sufficient for characterizing the risks, if any,
associated with exposure to total PAHs in fish (minimally) and shellfish (more likely)
caught and consumed from the HARS. The reason for this, in large part, is the lack or
absence of USEPA-promulgated toxicity criteria that could be used to characterize the
human health risks associated specifically with alkylated PAHs. In fact, although several
parent PAHs have been classified as probable human carcinogens (Group B2),
benzo[a]pyrene is the only parent PAH for which and oral cancer slope factor is currently
available (USEPA 2002).
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At present, the USEPA only recommends testing for seven of the listed parent PAHs
(benzo[a]pyrene, benz[a]anthracene, benzo[b]fluoranthene, benzo[k]fluoranthene,
chrysene, dibenz[a,h]anthracene, and indeno[l,2,3-cd]pyrene), which, with the exception
of benzo[a]pyrene, are classified as probable carcinogens (USEPA, 1999). USEPA also
recommends the use of relative potency factors for 14 PAHs (relative to benzo[a]pyrene)
in guidance for performing quantitative risk assessment of PAHs (USEPA, 1993). This
approach is recommended in USEPA guidance for evaluating chemical contamination in
fish (USEPA, 2000) to calculate potency equivalency concentration in fish tissue samples
for comparison to acceptable risk-based concentrations calculated for benzo[a]pyrene.
Thus, it does not appear that the inclusion of alkylated PAHs in risk characterizations is
warranted at the present time.
As indicated above, due to the apparent lack of compound specific toxicity criteria for
alkylated PAHs, measurement of alkylated PAHs would only significantly improve the
risk assessment of total PAHs if some sort of potency equivalency factor could be applied
to each of the 30 alkyl PAH homologues listed in Table 2 (of the peer review document).
One way to do this would be to apply the toxicity criteria associated with each parent
PAH to each of its alkyl homologues. However, in the absence of specific potency
testing the resulting uncertainty would be great. For example, from the information
provided in Table 2, it appears that most of the 30 alkylated homologues are associated
with noncarcinogenic parent PAHs (benz[a]anthracene and chrysene are only parent
PAHs with associated alkylated homologues)
Including alkylated PAHs in the HARS analytes list would likely only marginally
improve the risk assessment of non-cancer endpoints associated with PAHs, since cancer
risks for PAHs are generally associated with lower exposure concentrations than for
noncarcinogenic effects. However, several important aspects of the toxicology of PAHs
remain unknown. For example, alkylated forms of parent PAHs could actually be more
potent carcinogens than the parent PAH, and/or alkylated forms of noncarcinogenic
parent PAHs may actually be carcinogens. Benzo[a]anthracene is considered to be a
weak carcinogen, but the substitution of methyl groups on specific carbons of the ring
results in 7,12-dimethylbenz[a]anthracene, which is one of the most powerful PAH
carcinogens known (Cassarett and Doull, 1991).
' l\ the piopa\LiI adaptation of I PI Wilhod b2"t) (Ippindu /*/ aucpttihlc ana appropiuiTt lor
regulatory decision-making? If not, what isan acceptable andjappropriatejnethqd?t
As long as the selected analytical method can detect concentrations that are at or
preferably below concentrations associated with acceptable risk levels (i.e., risk-based
detection limit), the proposed method may be suitable for the proposed HARS
Framework. For example, in soils, a benzo(a)pyrene concentration of 0.062 mg/kg (62
ug/kg) corresponds to a risk level of 1.0 x 10~6 risk assuming residential exposure
assumptions for exposure via inhalation, soil ingestion, and dermal contact (USEPA
2000). Since the achievable detection limits associated with the USEPA Method 8270
are 10 ug/kg and 5 ug/kg for sediments and biological tissues, respectively, it would
appear that the proposed method might be acceptable the proposed HARS Framework.
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However, it should be noted that sample specific detection limits are typically required by
regulatory agencies when data are to be used for calculating potential exposure
concentrations. Thus, even though the method detection limits (MDLs) appear to be low
enough, the actual sample specific detection limits could be much higher if analytical
interferences (e.g., total petroleum hydrocarbons) are commonly associated with either
sediment or tissue samples.
I mder what specific conditioi
As noted above, due to a lack of established toxicity criteria for these compounds, it does
not appear that including alkylated PAHs would enhance the characterization of human
health risks associated with exposure to PAHs. However, it is not uncommon for 50% or
more of total PAH found in the environment to be comprised of alkylated PAHs. Thus, it
may be useful to require bulk sediment chemistry testing to ascertain levels in project
sediments to develop a better quantitative understanding of the composition of PAHs and
whether bioaccumulation testing is warranted. This type of information would be very
useful in the event that human or ecological toxicity criteria are developed in the future
for these compounds. The information gathered over time may even prove to be useful in
determining whether or not it is ever necessary to develop toxicity criteria for these
compounds.
5 W hut uncertainties would be introduced within the analyst*. o! risk should alkylated Ps,
indiuLd ' It hut \tep\ could be taktn to ulloupI for thf \e uriLttta-ntus in decision-mi
Clivcn the likelihood the method for using non-deiects (as described in EPA 'CENA N, I J>
As noted in charge #2, it appears that there would be a quite a bit of uncertainty
associated with both exposure and the toxicity of the alkylated PAHs. If this issue cannot
be addressed effectively, either through the development of toxicity criteria and/or a
potency-weighting scheme, the level of uncertainty associated with any risk
characterization of the alkylated PAHs would be so large as to potentially render the
information meaningless. Additionally, as noted above, it does not appear that inclusion
of the 30 alkylated PAHs would do much to enhance evaluations of potential cancer risks
associated with human exposure to PAHs, which is usually the main concern when
dealing with these compounds in a human health risk assessment. It would appear that
the uncertainty and increased cost associated with the inclusion of these compounds
could potentially outweigh any potential benefits pertaining to human health risk
assessment.
With regard to assumptions concerning detection limits, using the full detection limit for
non-detect measurements rather than assuming a concentration equal to one half of the
associated detection limit as recommended in risk assessment guidelines (USEPA, 1989)
when calculating exposure concentrations or comparing data distributions would likely
result in significant overestimations of risk. An implication of overestimated risks
associated with the use of full detection limits would be the erroneous failure of project
sediments in the risk assessment. Additionally, unless the same assumption is used to
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evaluate project sediments and reference / background areas, it is unlikely that statistical
comparisons of PAH distributions in potentially contaminated sediments to distributions
of PAHs representative of ambient PAH levels in reference area sediments would be of
any value (particularly if there were a large number of non-detect values associated with
either data set). The greater the number of non-detect values, the greater the chances that
a statistical comparison would erroneously conclude that the PAH distribution in project
sediments were higher than ambient (or reference) levels.
B2 Proposed Additions to Analvte List: Organotins
6 li is recognized that additional methods have heen used for the analysis of organotins (e.g.. Krone
i't al, f'MVJ Will the propmed analytical method (Rice rt alIV87) provide adequate data of
Sufficient quality to assess relevant risks from organotins? If not, please provide
recommendationsJ
Before the Organotin Antifouling Paints Control Act of 1988, tributyl tin (TBT) was
widely used in marine paints to prevent the growth of antifouling organisms such as
barnacles and mussels on the hulls of boats and ships (USEPA, 1997). Today, TBT is
widely recognized as an extensive contaminant in marine sediments and has been found
in high concentrations in fish and shellfish (Cal/EPA, 1999). TBT concentrations in
marine fish have been found to vary widely (see e.g., Harino et al., 2000, Yamamoto,
1994). In the United States, TBT water concentrations of 0.02 to 0.84 |ig/L in the Great
Lakes, 1.0 |ag/L in the San Diego Bay, and concentrations up to 0.8 |ag/L along the East
Coast have been reported (ETN, 1996). Because of TBT's relatively high toxicity to
marine invertebrates and its ability to accumulate in fish and shellfish, USEPA has
recommended a national water quality criterion of 0.01 jo.g/L for TBT in saltwater
(USEPA, 1999).
Significant differences have been found in the occurrence of different organotin
compounds (notably, mono-, di-, methlytin, butyltin, and octyltin) in the environment
(Cal/EPA, 1999; EC, 1993). Marine organisms, especially invertebrates, are extremely
sensitive to TBT and other organotin compounds. Adverse effects on marine organisms
include reproductive toxicity, immune system dysfunction, nervous system disorders, and
mortality (ETN, 1996, US ACE, 1998). However, factors such as organic carbon in
sediment and water, pH, salinity, clay fraction, and the presence of inorganic constituents
such as iron oxides have been shown to affect the toxicity and bioaccumulation of TBT in
marine organisms (USEPA, 1999).
Several analytical methods for measuring TBT can be found in the scientific literature.
Comparing the analytical method of Rice et al. (1987) to that proposed by Krone et al.,
(1989), it would appear that the Krone et al. (1989) may be the preferable analytical tool
for measuring all of the butyl-tin species. Krone et al. (1989) appears to addresses the
important analytical considerations necessary to accurately assess the extent and impact
of TBT-related contamination in the marine environment. These considerations include
sample storage and preparation, accuracy and precision of analytical methods, reporting
convention, and quantitation method.
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ited taeiisure thequalily and usability of the
Sample storage and preparation are important factors that are often overlooked.
Although speciation changes or losses of organotins during storage are not well
established, a review of analytical procedures by Abalos et al. (1997) indicates that
freezing biological and sediment samples preserved the stability of organotins in the
sample matrix for at least 3 months. Some studies indicated that organotins were mainly
associated with the fine silt and clay fractions of sediments rather than the larger sand
fractions. In studies where sediments were sieved, only fractions below 100 |im were
analyzed. Although the effects of drying samples are not well known, studies have
indicated that oven-drying samples at high temperature have resulted changes in tin
speciation. In some cases, dried biota sample produced lower recoveries due to
decreased extractability of organotin rather than organotin degradation. (Abalos et al.,
1997)
Accuracy and precision of an analytical method should be monitored using calibration
standards and spike samples. Although the lack of certified reference materials and
derivatized organotin calibrants for gas chromatography may limit the ability to measure
TBT recovery rates, the use of multiple spike samples at concentrations typically found in
the environment has been suggested (Abalos et al., 1997). More detailed information
about specific testing methods is found in Abalos et al. (1997).
A related and important consideration is reporting conventions. Historical data have been
lacking continuity, resulting in confusion when comparing TBT data from different field
and experimental studies. According to US ACE (1998), TBT as Sn, TBT, TBTC1, or
TBTO represent different species of TBT, the numeric values calculated from these data
can be confusing to interpret. US ACE (1998) suggests that all data be reported in
accordance with current reporting practices used in the scientific community.
8. Under what specific conditio/is would tk e testing for orgimuiins for a particular project be
Currently, bioassay testing is required for samples exceeding a TBT criterion threshold in
bulk sediment. US ACE (1998) specifies bioassays include 10-day mortality test,
sediment larval bioassay, and a 20-day biomass test. Unfortunately, recent studies have
shown that the majority of these tests may not evaluate TBT toxicity appropriately.
Serious acute and chronic toxicity in the bioassay organisms appears to necessitate longer
exposure periods than specified in the standard bioassay protocols.
The guidance in US ACE (1998) appears to be an appropriate basis for including or not
including consideration of organotin compounds in the HARS Framework. According to
USACE's (1998) review of TBT sediment and biota data, sediment chemistry screening
levels and bioassay testing are both poor indicators of the actual environmental effects.
Instead, USACE recommends direct measurements of TBT concentrations in interstitial
water and tissue are better methods of assessing TBT toxicity in marine environments.
Specific testing and report methods are found in USACE (1998).
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B3. Proposed Additions to Analvte List; Coplanar PCB Congeners
9._ If the approach Jar evaluating dioxin ;.s mat.
For the purposes of human health and ecological risk assessment, the approach to
evaluating dioxins should include co-planar PCBs. Sediment chemistry and
bioaccumulation results should be reported both as total toxic equivalents (TEQs) and as
individual chemical measurements. It is generally recognized in the scientific literature
that 13 PCBs with a co-planar structural configuration elicit toxic effects in mammals,
birds, aquatic animals through an Ah-receptor mediated mechanisms similar to the
2,3,7,8-substituted dioxins and furans (van den Berg et al., 1998). In Canada (2001) and
in the U.S. (USEPA, 2000a), environmental quality values and risk assessments include
consideration of both 2,3,7,8-substituited PCDD/Fs and co-planar PCBs and assume the
effects in humans and wildlife are additive. The results of the USEPA dioxin
reassessment are likely to indicate that the dioxin-like should be included in both health
and ecological risk assessments (USEPA, 2000b). The approach for risk assessment
involves the application of World Health Organization toxic equivalent factors (TEFs)
(van den Berg et al., 1998).
C. Comparison to Reference
The proposed HARS Framework indicates that the concentrations of contaminants in
tissues of organisms exposed for 28 days to project sediments will be compared to the
concentrations of contaminants in tissues of organisms similarly exposed to reference
sediments. The reference sediment is characterized as a clean, sandy material collected
from an area in the New York Bight located south of the HARS, where sediments are
assumed to be unaffected by past or future dredged material disposal activities.
In accordance with the Green Book, the HARS Framework specifies that undetected
residues ("< detection limit") in bioaccumulation testing of project sediments will be
assumed to as Vt the detection limit if below the method detection limit; if the detection
limit is above the method detection limit, the residue concentration in project sediment
tests will be assumed as the reported detection limit. However, for the reference
sediment tests, the residue concentration will be assumed as zero. Project sediments are
suitable as remediation material if the results of 28-day bioaccumulation testing are
statistically equal to or below (at a 95% confidence limit) the results of reference
sediment tests. For chemical classes of compounds such as PCBs, PAHs, DDT, and
endosulfan, results will be evaluated based on the total residue concentration of that class
of chemicals.
The approach described in the HARS Framework should be reconsidered for three
reasons:
a. The physical composition of the reference area sediment (clean, sandy material) is
inappropriate for use as background, or uncontaminated material;
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b. The strategy for addressing detection limits does not appear to include statistical
considerations such as the likely distribution of residue concentrations;
c. Analysis of functional groups of chemicals (or, chemical classes) rather than
individual chemicals within a class is not appropriate for risk assessment
evaluation.
With regard to the physical characteristics of reference sediment, comparison of
bioaccumulation test results from a fine-grained, organic rich material with results from a
coarse-grained, sandy, organic poor material appears to be inappropriate. According to
USACE-NYD (1992), reference sediment should be substantially free of contaminants,
but similar in grain-size distribution, organic content, and % moisture to the proposed
dredged material and reflective as possible of hydrographic conditions characteristic of
the disposal site. The current reference area does not meet the latter requirements.
Among the possible options, one approach is to define several reference sites where the
physical characteristics (namely, grain-size and organic carbon content) of the sediment
closely match the characteristics of sediment generally found throughout NY/NJ Harbor.
Bioaccumulation testing involving a suite of reference sediments from several locations
might be preferable to reliance on a single location because testing and statistical
comparisons could capture the natural variability of materials with varying grain-size and
organic carbon content.
IU. Please consider the policy for ussigning values tat one half the detection limit) to tissue residues,
that are reported us " ''.detect ion limit" (destrihed in EPA'CblNA N, IW7} as you r-view the
proffostd evaluation methodologies As you deem appropriate, please eommt nt on the ejfec *\ <
thi\ pobc\ outfit outlined t\aluation\ If the i urrt nt t.ppt oath s lonstdettdinuppropnute, vvA
would be a technic ally \uppartable alternative a,
detection limit" in the risk assessmentpraces.
The approach to addressing non-detect values in an environmental data set is a classic
challenge for both human health and ecological risk assessment. Several statistical
routines have been proposed to censure data sets or to evaluate and select the appropriate
option for a given data set (USEPA, 2000a, 2000b). The important concern here is that
reliance on detection limits could create a false sense of contamination or toxicity,
particularly if laboratory detection limits are high. The upshot is that decisions could be
influenced more by the competency of the analytical laboratory and less by the actual
condition of the project sediment.
For evaluating samples that are reported as "
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non-detect values in environmental samples - and often the least satisfactory - is to
assume that the chemical is present is at one-half the detection limit (USEPA, 1989).
In recent guidance, USEPA (1997, 2000b) recommends several alternative statistical
methods for evaluating a sample population that includes both non-detect and positively
detected values. Depending on the magnitude and frequency of non-detect samples,
different statistical evaluations are suggested (USEPA, 2000a, 2000b). For data sets with
a small number of non-detect samples (i.e. <15%), USEPA (2000a) generally considers
using one half the detection limit for non-detect samples to be a reasonable estimate of
risk and exposure. To more accurately estimate mean chemical concentrations, the
Cohen's method, Trimmed Mean, or Winsorized Mean and Standard Deviation are
suggested for data sets with 15-50% of non-detect samples. For data sets containing
>50% non-detect values, USEPA guidance recommends statistical tests of proportions to
test hypotheses using the data. (USEPA, 2000a, 2000b) The statistical methods
mentioned above are fully described in USEPA (2000b).
11. . /a the use offunctional gioupmgs in \tutisucal companions to rcjirsnLe appropriate a nil'or
ruble to
A second long-standing concern to both human health and ecological risk assessment
concerns the appropriate methods for evaluating a class of chemicals such as the dioxins,
PCBs, PAHs, and DDT. In general, it is always preferable in a risk assessment to
evaluate individual chemicals, rather than rely on analysis of functional groups of
chemicals based on total concentration. This concern is particularly relevant to sediment
assessments and the proposed HARS Framework.
With regard to PCBs, available USEPA potency estimates are predicated on total PCBs
and selected Aroclor formulations and offer little insight into the behavior and toxicity of
the most important PCB congeners on human health and wildlife. Most scientists now
agree that future PCB risk assessments should evaluate non-ortho- and selected mono-
ortho- PCBs, and consideration of exposure models and studies involving specific PCB
congeners alone or in combination with the dioxins and furans (NRC, 2001; Eisler,
2000). For example, Coates and Elzermann (1986) demonstrated that equilibration times
for PCBs in sediments varied considerably from a few weeks for PCBs with low chlorine
content to months or years for PCBs with significantly higher chlorine content. Lick and
Rapaka (1996) observed that slow rates of adsorption / desorption may significantly
modify the level of toxicity inferred from measurements of the chemical concentration in
the sediment, particularly if equilibrium partitioning is assumed but the chemical in the
pore water and colloidal phase is not in equilibrium with the chemical sorbed to the solid
phase. This will be a potentially important concern for aged PCBs in sediment, where
degradation or a change in the congener profile relative to the initial discharge may have
changed over substantially over time.
USEPA acknowledges that potency factors generally developed from laboratory studies
involving mixtures, may not be appropriate in situations where the mixture has aged and
the distribution of congeners in the mixture have changed (USEPA, 2000). In fact, the
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Agency identifies two different potency factors for PCBs, one based on an
immunotoxicological endpoint for exposure to Aroclor 1254, and the other based on a
reproductive endpoint for Aroclor 1016. Given that both PCB formulations contain
different distributions of individual PCB congeners which may change due to weathering
processes (Jarman et al., 1997) and that the congener distribution in project sediment can
only be determined by congener-specific bulk sediment chemistry analysis, it would
appear to be appropriate to rely on congener-specific measurements to ascertain which
potency factor is most appropriate for risk assessment.
With regard to PAHs, the approach most often advocated for risk assessment is a toxic
units model similar to that first proposed by Swartz et al. (1995). The approach has
several limitations and uncertainties such as differences in mechanism of toxicity,
bioaccumulation potential, degradation rates, and occurrence. These uncertainties were
noted by Johnson (2000) in a study relating PAH levels in sediment and effects in winter
flounder. Nonetheless, similar to the PCBs, it is more appropriate in risk assessment to
evaluate the individual PAH compounds rather than rely on the potency of one
component within a mixture or functional group. The state of the science is sufficient for
evaluating exposure to the analytes listed in Table 2 of the proposed HARS Framework.
D. Adjustments to Steady State
Dl. Adjustment to Steady State; Organic compounds
12. Is it appropriate to apply u multiplier based on log KUiifoi these compounds (or %unii s,
thera other specific data thai can be used to estimate steady state? If&o, please idemi,
H ij/ii'M the h\drophobu.it\ ojalkyluhdI'AHs, is the use oj tne lorrec turn J( i tor
associated with the corresponding parent an appropriate approach for estimating st*uly stu
residues of alkylated Pillsr> If not. pit' ive eltihni ate
14. For the DDT derivatives and dieldrin, pleasa comment on the appropriateness of using ,W. n
data}uthci than S urens-speufie data in the estimation pj steady state multipliers
15 Irethe ippioaihes taken to adjust organic contaminant hioanumulation data to suadi s'a
adequate9 Do the proposed multiplier.'- agree -.villi previously published stud ts(u do they
Whai are the major sotutes of'uicertainh assmiated with the appioai lies'' What aliernativ
approaches would reduce the i
D2. Adjustment to Steady State: Metals
/ 7. In your opinion, is the methodology1 followed to derivt the steady state multiplier Joi non-esst nt
metals (i e, ajucloi of thtee) suuitifiiallj appropriate fApptndix (j) ' Pleasa elaboiute
na\t anv i n ommendations of additional or al emote mil
used to either \upplrmimpr replace tl.rproposed methvi
E.
Human Health Evaluation
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El Human Health Evaluation: Overall
Risk assessment is the final step in the proposed HARS Framework (Figure 2). The
satisfactory results of a risk-based evaluation is the final decision point, presumably, if
the results of liquid, suspended particulate, and solid phase testing do not exceed
prescribed threshold criteria, and if the results of bioaccumulation testing do not differ
significantly from similar testing using reference sediment, and if tissue residues of Table
1 analytes from the 28-day bioaccumulation testing do not exceed regional matrix values,
dioxin matrix values, FDA action levels, or ambient fish tissue levels at the HARS when
compared to predicted tissue concentrations in upper trophic level fish using a food web
model.
There are two aspects to the risk assessment evaluation. One aspect involves calculation
of HARS-Specific values for different contaminant residues using a backward calculation
approach to determine threshold concentrations in worm and clam tissues for comparison
to the results of the 28-day bioaccumulation test (corrected, if necessary, to steady-state).
The other aspect involves a forward calculation approach to determine whether the risks
to anglers consuming fish caught at the HARS exceed specified cancer and non-cancer
thresholds.
As illustrated in Appendix B, the risk-based evaluation includes four components: (a)
consideration of steady state bioaccumulation and food web transfers, (b) comparison to
background tissue concentrations, (c) consideration of ecological effects, and (d)
consideration of cancer and non-cancer effects on human health. Human health risks are
presumed possible only through consumption of seafood caught from the HARS. A
simplified food web model comprised of three trophic levels is used to estimate
contaminant migration from sediment to benthic organisms to benthic predators to upper
trophic level fish potentially consumed by recreational anglers. Several assumptions in a
deterministic exposure equation are specified to predict angler fishing behavior, angler
fishing success, seafood consumption, and residue losses by cooking. Cancer and non-
cancer risks are evaluated using USEPA potency factors established for the Superfund
program. The approach used to evaluate the health risks associated with lead is generally
consistent with USEPA's biokinetic model.
18. Please c tmmciit on each factor listed tihuw (and in J.ihL 5) as to its appropriateness for use in
[lit. <.qua!ion\ h\udako\e Huuld)uu rnonmendudditionJJut.U»i Mould iou <.hangi or
modify tht equations wntten «/)ou ' IJ mi, ho\i *
19 ¦ire the methods usi d to den\e tht' human health e\po\ari puntnu ier\ and ti\\ii>ned valuer
Cancer potency factor/reference dose: The use of potency estimates developed by
USEPA for the Superfund program is suitable for risk assessment. The Agency is
encouraged to allow consideration of new peer-reviewed toxicological information that
may result in a potency estimate (cancer and/or non-cancer) different from that identified
in IRIS. The burden of scientific evidence should be placed on the proponent and
discussed m Slliioh L approp'iatt (plea\i nu m the reft rt.ni id appending)' Hnot pU
elaborate Hon should the\t jut tors he Jul tor,d into the risk anulyu \ und deusion-nuikh
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evaluated on a case-by-case basis, or until agreement is evident within the scientific
community.
Seafood consumption: The population of high consumers (defined as New Jersey
recreational anglers) who preferentially fish at the HARS and, presumably, obtain all of
the recreationally caught seafood in their diet from fishing at the HARS are assumed to
consume 7.2 grams per day (g/day) fish. According to USEPA, the consumption rate of
7.2 g/day is intended to be a "site-specific estimate of daily fish consumption and was
derived from a fish consumption surveys conducted by New Jersey Marine Sciences
Consortium (NJMSC; 1994).
While there may be ethnic groups in the NY/NJ metropolitan area who consume large
amounts of fish in their diet, there is no evidence to indicate that a population (of any
size) of recreationally anglers obtains all of the recreationally caught seafood in their diet
only from the HARS. As such, the factor used by USEPA is an unreasonable estimate of
a "high end" fish consumption rate. According to NJMSC (1994), the average
consumption rate for recreational fish (freshwater and marine combined) may be lower,
approximately 4.5 g/day among New Jersey anglers. Clearly, subsistence fishing does
not occur at the HARS. And it is difficult to imagine that a fish consumer within the
NY/NJ metropolitan region purchases and consumes seafood commercially caught at the
HARS in amounts that would approach those by recreational anglers fishing at the
HARS.
Several studies suggest that recreationally caught seafood consumption rates could be
even lower, and as low as 1.0 g/day (Ebert et al., 1994; Ruffle et al., 1994; Price et al.,
1994; USEPA, 1999). According to USEPA (1999) Exposure Factors Handbook, the
average daily intake of marine fish in the Atlantic region is 5.6 g/day using National
Marine Fisheries Service survey results from 1993. The available literature clearly
indicates a wide range of assumption depending upon sex, age, species of fish consumed
and region of the country. A wide range of views on this subject is evident within the
RMW from the materials included in the peer-review document.
Another important consideration not evident in the proposed methods for health risk
evaluation is consideration of different populations of potentially exposed individuals,
namely prospective parents, pregnant women, and children. However, rather than adopt
different point estimates of exposure as part of deterministic analysis of these potentially
sensitive receptor groups, a probabilistic approach that that incorporates a wide range of
consumption habits (as well as those associated with other exposure assumptions) would
be an appropriate approach for addressing the potential for exposures to specific
populations of fish consumers.
The uncertainties associated with this assumption are simply too great and the available
information too outdated to be used for risk assessment. The importance of this
assumption is clearly evident in von Stackelberg et al. (2002). The Agency is strongly
urged to undertake a fish consumption survey of the HARS targeted at developing a
scientifically defensible understanding of recreational angler fishing habits, their catch,
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and consumption habits. Such a survey should consider guidance from several sources to
minimize the various sources of uncertainty that can adversely affect the quality of the
information generated from interview, observation, or self-reporting survey methods
(USEPA, 1992; Ebert et al., 1994; USEPA, 1995).
Exposure duration: The assumption of 70 years is consistent with traditional risk
assessment practices. Data on residence time in USEPA (1999) indicates that the average
individual (whom may fish recreationally at the HARS) is likely to live in the same area,
on average, for only 9 years (50th percentile) and 30 years (95th percentile). With regard
to the average human lifetime, current data suggest that 75 years would be an appropriate
value to reflect the average life expectancy of women (78.8 years) and men (72.2 years)
in the general population (USEPA, 1999).
Site use factor: The site use factor (SUF) factor is intended by USEPA to represent the
proportion of time that fish caught and consumed by anglers at the HARS are exposed to
contaminants in sediment directly or through feeding on benthic prey residing at the
HARS. According to Appendix J, the value of the SUF (0.777) was estimated using 1993
data on commercial fish landings and assumptions regarding the proportion of different
fish to the typical angler's total fish diet to determine the weighted seasonal residence
time of fish caught and consumed by anglers at the HARS. Precisely how the Agency
performed the calculation could not be ascertained by this reviewer.
The use of an SUF is appropriate for risk assessment. The significance of this factor is
clearly evident in Linkov et al. (2002). It is inconsistent with the current understanding
of the life histories and environmental requirements of several coastal fishes and
invertebrates (see e.g., NMFS, 1999a, 1999b). The assumption that both lower and upper
trophic level fish caught by recreational anglers at the HARS inhabit and forage
exclusively at the HARS for their entire life history is overly conservative. The
migratory and foraging behaviors of individual species and seasonal variations in the
presence/absence of different fish populations clearly suggests that continuous lifetime
exposure and accumulation of contaminants originating in sediment is implausible.
For example, USEPA assumes nearly half of the fish caught and consumed by anglers in
the HARS are winter, summer, or yellowtail flounder. According to NMFS (1999a,
1999b) and USFWS (1989a), winter, summer, and yellow flounder the behavioral
characteristics (including foraging, habitat use, and migratory patterns) are different for
each of the three species. Migratory habits of flounder suggest that these fish spend less
than 6 months, on average, in coastal waters that may, or may not, include the HARS.
Similarly for bluefish, USFWS (1989b) profile of life history and environmental
requirements indicates a considerable migratory and foraging ranging, which is indicated
in the 1993 commercial landings data by the relatively small or zero catch volume
reported between October and March.
Aside from fish life history, commercial catch data from 1993 are unlikely to be
representative of the current condition of the mid-Atlantic fishery. NMFS fisheries
scientists and northeastern fishery councils suggest that several recreational and
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commercial fish stocks have declined over the past 5-10 years (NMFS, 2000). Similarly,
the proportion of different fish species consumed by anglers also is based on nearly
decade-old information developed by NJMSC (1994), and may not reflect current fishing
success at the HARS.
In addition, the extrapolation of 1993 commercial landings data for the entire New York
Bight to the HARS appears overly conservative. The HARS encompasses approximately
16 square miles, or approximately less than 0.1%, of the 19,000 square miles of fish
habitat in the New York Bight. This significant difference in geographical area would
suggest that estimates of the fish catch attributed to the HARS are likely to be grossly
inaccurate.
In summary, the approach developed by USEPA to quantify an SUF (aside from this
reviewer's inability to replicate the calculation) is flawed and requires reconsideration
(notwithstanding that the calculation logic described in Appendix J is almost impossible
to understand). The revised approach described by Pavlou and Shephard (2000) should
be considered in the re-evaluation of this factor. Pavlou and Shephard (2000) appear to
have considered the mobility of target fish, the proportion of the HARS included in the
New York Bight, and the proportion of the year that a given species might forage or
inhabit the HARS. The Agency also should carefully consider the approach described by
Linkov et al. (2002), which illustrates nicely the use of probabilistic methods and the
significance of assumptions on habitat size, forage and habitat attraction to target fish
species, as well as other factors on human health risk assessment.
Trophic transfer factor: The use of trophic transfer factors (TFF) for chlorinated organic
chemicals, metals, and PAHs is intended to estimate the amount of contaminant that
transfers from benthic prey to fish predators in USEPA's three-compartment food web
model. The predicted body burdens in upper level predators (e.g., bluefish, cod, and
striped bass) represents the exposure point concentration for recreational anglers catching
and consuming fish from the HARS. It is also a very important consideration in the
calculation of HARS-Specific values.
The use of TFFs is appropriate for risk assessment. It is well established that
contaminants in sediment have the potential under certain conditions to migrate into biota
and then undergo one or more transfers to other higher trophic level organisms in a food
web. This may result in higher concentrations accumulating in upper level predators,
which may adversely affect the organism or pose an adverse health risk to anglers who
catch and consume upper level predators. In the absence of spending considerable
resources measuring each stage of a food web, several frameworks have been developed
to predict contaminant transfer rates and associated concentrations of different classes of
contaminants in different organisms and/or trophic compartments. The Gobas et al.
(1993) food web model is among the approaches often identified by ecological risk
assessors to evaluate chlorinated organic contaminants and represents a reasonable
approach by USEPA for this class of chemicals. A bioenergetics model is another
approach that has been advocated in ecological modeling studies conducted in the NY-NJ
Harbor area (Iannuzzi et al., 1996; Thomann, et al., 1992; Thomann, 1989).
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The approach used to determine TFFs for PAHs, mercury, arsenic, and metals was not
evaluated.
The Agency should re-evaluate the proposed TFF of 3 for PCBs, DDT and, perhaps,
other chlorinated organic chemicals. Because the Gobas et al. (1993) food web model is
sensitive to the contaminant's octanol-water coefficient (Kow), TFFs for some congeners
within a functional class of chemicals may be over- or under- estimated. For example,
the TFF of 3 may be appropriate for tetra- to hexa- chlorinated PCBs, but not for
congeners with fewer than 4 chlorines or more than 6 chlorines. This may be important if
the results of sediment chemistry analysis indicate the presence or absence of different
PCB congeners that may have higher or lower propensities to bioaccumulate in a food
web.
In their review of a framework for evaluating bioaccumulation in food webs, Sharpe and
Mackay (2000) identified several general model limitations that are relevant to the
proposed HARS Framework. Among the important factors noted by Sharpe and Mackay
(2000) that are rarely considered in food web models: natural variations between
organisms of the same type, diet and physiological changes at different life stages,
chemical distribution within an organism, and depuration time between uptake by prey
and consumption by predators. In the absence of a complete or satisfactory
understanding of these and other factors, probabilistic exposure modeling similar to that
proposed by von Stackelberg et al. (2002) represents the best available approach to
predicting the trophic transfer of contaminants and associated risks to humans.
20 It :hi.' approiith taken to i elate fi\n whuh- body and fdlei cunt t ntiat oris u lenliju ullv
Appropriate? If not, what method would you recommend?,
Appendix K of the proposed HARS Framework specifies different whole body to fillet
factors (BFR) for organic lipophilic chemicals (1.35 for PAHs, PCBs, pesticides, and
butyl tins), arsenic (1.4), chromium (1.2), mercury (0.7), and other metals (1.0). The
technical information supporting the BFR for organic lipophilic chemicals is not provided
in the document. For metals, BFRs were adopted from Bevelhimer et al. (1997). The
BFR is particularly important for derivation of HARS-Specific values. The approach
specified in the proposed HARS Framework appears to be reasonable.
21. _ u Could the analysis be unproved by focusing on key Jish (s> eafood)jpecies at the OARS? What^
According to Appendix I, USEPA indicates that it conservatively assumed that all fish
caught and consumed by recreational anglers would consist entirely of species that
inhabit the HARS. However, according to NJMSC (1994), among fish consumed by the
average New Jersey recreational angler, only approximately 53% are saltwater fish
(NJMSC, 1994). Furthermore, only approximately 32% of the fish consumed by
recreational anglers are marine species believed to inhabit or forage at the HARS. For
example, while USEPA suggests that flounder represents approximately one-half of fish
caught and consumed from the HARS, NJMSC (1994) reported that flounder comprises
only 18% of the saltwater fish consumed by recreational anglers (or 10% of the average
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total fish diet). USEPA also indicates that cod, bluefish and striped bass each comprise
approximately 11% of fish caught and consumed from the HARS. However, according
to NJMSC (1994), cod and bluefish each comprise only 4% and striped bass less than 1
% of marine fish consumed of the average recreational angler diet.
Indeed, the risk assessment approach described in the proposed HARS Framework could
be markedly improved by reliance on current HARS-specific information on species
inhabiting the HARS and the frequency of catch at different times of the year among the
population of recreational anglers that fish at the HARS. These three important pieces of
information - fish population, fish catch, and frequency of catch by anglers would greatly
improve the risk assessment.
22. In your opinion, is the approach for assuming lata! metal to be in the most toxic form appropriate
and reasonable? Should metal speciationJcomplexation he considert d in the a\.\o.
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risk assessment of total PAHs bioaccumulated if some sort of potency equivalency factor
could be applied to each of the 30 alkyl PAH homologues listed in Table 2. One way to
do this would be to apply the toxicity criteria associated with each parent PAH to each of
its alkyl homologues. Doing so here, as part of the proposed HARS Framework, is not an
unreasonable approach; however, it is unclear whether the approach will significantly
improve the human health risk assessment.
Please comi
It is not unreasonable to include the consideration of PAHs in a fish consumption risk
assessment. In some cases, the health risks associated with exposure to PAHs through
consumption of contaminated fish can be significant, on the order of the risks posed by
metals and other contaminants. For example, in the Galveston Bay National Estuary
Program / Texas A&M seafood risk assessment (Brooks et al., 1992), the risks to seafood
consumers exposed to heavy metals, PAHs, pesticides, and PCBs in the edible portions of
oysters, blue crabs, spotted sea trout, black drum, and southern flounder were evaluated
to determine the probability of adverse health effects from eating contaminated seafood.
Most of the cancer risk was associated with PCB and PAH concentrations, with PCBs
providing a larger portion of the overall risk. Analysis of the specific PAH compounds
indicated that most of the PAHs originated from combustion products (such as engines,
burning, etc.) that were not directly associated with release of oil to the bay (Brooks et
al., 1992). In general, however, the risks to human health are generally believed to be
low.
What are the major sources nf uncertainty associated with the approaches de\a ibed in Sect inn
What alternative approaches would reduce the uncertainties? J low could tin \t una Ha ntiL s be
As indicated in response to charge #1, several resources are available describing the
numerous sources of uncertainty in human health risk assessment (e.g., US ACE, 1999;
USEPA, 1989; Paustenbach, 1995), sediment assessment (e.g., Ingersoll et al., 1997;
Dickson et al., 1987), and sediment bioaccumulation studies (e.g., Sharpe and Mackay,
2000). Among the most important uncertainties in the proposed HARS Framework are:
(a) contaminant levels in project sediment; (b) current environmental conditions at the
HARS; (c) relevance of test organisms in bioaccumulation tests to organisms that inhabit
the HARS; (d) food web transfers of contaminants in sediment to prey and to upper
trophic level fish caught and consumed by anglers; and, (e) fishing and consumption
habits among different angler populations.
The major sources of uncertainty specifically associated with the fish consumption risk
assessment approach described in Section E of the proposed HARS Framework are
typical of most deterministic human health risk assessments (see e.g., Wenning, 2002;
Paustenbach, 1995; USACE, 1999). Uncertainties fall into four categories consistent
with the four components of risk assessment- hazard assessment, dose-response
assessment, exposure assessment, and risk characterization. The uncertainties associated
with each of these components are discussed in detail elsewhere and should be reviewed
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and considered by USEPA (see also, e.g., USACE, 1999; USEPA, 1992b; Paustenbach,
1989).
The major sources of uncertainty that deserve especial consideration are as follows:
• First, the proposed HARS Framework appears to overemphasize the maximally
exposed individual (MEI) through repeated use of several conservative exposure
factors and contaminant fate and effects assumptions. Reliance on NJMSC (1994)
fish consumption surveys, 1993 commercial fish landings data, and the frequency of
fishing activity assumed by recreational anglers who fish at the HARS strongly
suggest that an atypical receptor population is evaluated in the risk assessment.
Current USEPA (1997) exposure assessment guidelines caution against such an
approach, in general, by noting that a worst-case or MEI analysis should be used only
to decide if an exposure is insignificant and should not be used to characterize the
actual or plausible human risks. Although the risk for sensitive human populations
should be understood, the typical levels of exposure for the majority of anglers should
be the focus of the risk assessment. To help minimize the potential for
misunderstanding, the USEPA should take steps to quantify the population of anglers
at the HARS. By characterizing the potentially exposed population, the risk
assessment could be modified to describe the number of exposed persons in different
angler subgroups, along with the most likely and upper estimates of exposure and the
associated plausible risk.
• Second, the risk assessment includes repeated use of several conservative
assumptions, which may dictate negative results for most, if not all, project
sediments. The values used to represent seafood consumption rates, site use factors,
trophic transfer factors, and exposure duration appear to be higher than warranted by
supporting information. Other assumptions implied in the risk assessment such as
fishing frequency, fishing success, fish behaviors, contaminant bioavailability, and
the food web also appear to represent the high end of most likely conditions.
• Third, the proposed approach does not account for the environmental fate of
chemicals. Many factors such as degradation by sunlight, soil and water microbes,
and evaporation can dramatically influence the degree of human exposure.
Sedimentation rates, geochemical changes that result in complexation are not
considered. The potential for dispersion of placed material over time is not
considered, though such processes are contemplated in the HARS SEIS. The
assessment approach assumes contaminant concentrations measured in project
sediments at the outset will remain constant and, presumably, bioavailable, in the
future.
• Fourth, the proposed approach to risk assessment doesn't include detailed guidance
on statistical analyses of environmental data. USEPA (1997) guidelines for exposure
assessment indicates that inappropriate statistical analysis of environmental data is
one of the obvious sources of uncertainty and most easily corrected of the common
errors in exposure analysis. This is specifically relevant to techniques for statistically
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handling samples having no detectable amount of contaminants. Several approaches
have been suggested such as using 50% of the limit of detection (LOD) to calculate
the plausible degree of human exposure when no detectable amount is identified;
others approaches include dividing the LOD by the square root of two, and other
much more complex approaches also have been recommended (Paustenbach, 1995).
• Fifth, uncertainties are likely to remain high because few of the exposure assumptions
and/or model estimates are validated using field measurements. As field
measurement techniques are further refined, the Agency should be in a position to
rely less on mathematical models for predicting chemical distributions in the
environment, with greater emphasis (and confidence) on actual environmental
information.
How can these uncertainties be overcome or managed? The problems associated with the
repeated use of overly conservative assumptions and the need to properly account for
small (but highly exposed) populations can be overcome by using a probabilistic
statistical method in the risk assessment. The probabilistic approach, which is typically
undertaken using a Monte Carlo statistical technique, can address the deficiencies evident
throughout the risk evaluation framework and elsewhere in the proposed HARS
Framework. It appears that most, if not all, of the various technical concerns identified in
this review and by the RMW reflect the Agency's reliance on a point estimate (i.e.,
deterministic) approach to the risk assessment.
A probabilistic approach is capable of incorporating the range of known or suspected
values for each environmental and exposure factor used in the exposure assessment
equation. Admittedly, these distributions will be dependent upon the availability of
credible information to characterize the range of behaviors, activities, concentrations,
etc., associated with each parameter. Nevertheless, instead of presenting a single point
estimate of risk, probabilistic analyses can characterize the range of potential exposures
and associated health risks and their likelihood of occurrence. In addition, the factors that
most affect the risk assessment results can be easily identified and resources assigned to
further understand these key factors.
Guidance is available from several USEPA sources (USEPA, 1999,1998,1997a, 1997b).
The Agency is strongly encouraged to consult these guidance documents and to adopt a
probabilistic approach as a first step towards addressing many of the technical issues
identified in this review.
26 HtuitiA\our uummLmlution }n> i\alutitinn the potential to vnt\ «/ or^unmns' SnuuU rn<. i bi
evaluated ai individual compound^ Summed a\ total'' Should there be same eon\idcmtinn of
relutive torn ity'S
Organotin compounds should be included in the proposed HARS Framework and
evaluated in the risk assessment. The individual compounds should be examined in the
risk assessment, because potency estimates and occurrence in sediments may vary
considerably in project sediments. In the absence of chemical-specific potency estimates,
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a relative potency scheme should be adopted, if such an approach is evident in the
scientific literature.
27. . Plcmr comment on the appropriateness of the proposed approach Jrrconvcrti.
analytical data for alkylated and parent PAHs to estimate risk Jwm all PAHs.
This question has been addressed elsewhere in this review.
E2. Human Health Evaluation: Comparison to HARS-Specific Values
28. Do you believe that the "disaggregate"! mi/deling discussed above (and shown in Figure 4) far
estimating human health HARS-Specific Values for lead is appropriate? Would you r*cor,
an alternative tLk asses
JO. Js the conceptual model for evaluatingfish i
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exposure, through ingestion of seafood appropriate and reasonable? How can the uncertainties;
associated with the assumptions in this conceptual model be reduced/ Pleoie coniider the spatial
and temporal elements of exposure in vour discussion \
The approach to calculating cancer and non-cancer health risks is reasonable.
The conceptual model for evaluating fish exposure to dredged material at the HARS is
reasonable. The conceptual model for evaluating human exposure through seafood
consumption is reasonable. However, in most cases, the information used to support
some of the exposure assumptions is either insufficient or inadequate to support the
values assigned to different exposure assumptions. A significant portion of the
uncertainties associated with the proposed HARS Framework could be resolved either
through (1) field validation of key assumptions such as angler activity at the HARS and
frequency of catch, and (2) use of a probabilistic, rather than deterministic, analysis of
food web transfers and human exposure to contaminants in project sediments.
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V. REFERENCES CITED
Section III References:
American Fisheries Society (AFS). 2000. Proceedings of the AFS Forum on
Contaminants in Fish, October 18-20, 1999. Bethesda, MD. Prepared by EVS
Consultants. August.
Ankley, G.T. 1997. Laboratory vs. field measurement endpoints: a contaminated
sediment perspective. In: Ingersoll, C.G., Dillon, T., Biddinger, G.R. (Eds.). Ecological
Risk Assessment of Contaminated Sediments. SET AC Press. Pensacola, FL. pp. 115-122.
Chapman, P.M. 1989. Current approaches to developing sediment quality criteria.
Environ. Toxicol. Chem. 8: 589-599.
Chapman, P.M., Wang, F. 2001. Assessing sediment contamination in estuaries. Environ
Toxicol Chem 20: 3-22.
DelValls, T.A., Foija, J.M., Gomez-Parra, A. 2002. Seasonality of contamination,
toxicity, and quality values in sediments from littoral ecosystems in the Gulf of Cadiz
(SW Spain). Chemosphere 46:1033-1043.
Dickson, K.L., Maki, A.W., Brungs, W.A. 1987. Fate and Effects of Sediment-Bound
Chemicals in Aquatic Environments. Pergamon Press, New York, NY. pp. 449.
Engler, R.M. 1980. Prediction of pollution potential through geochemical and biological
procedures: Development of regulation guidelines and criteria for the discharge of
dredged and fill material. In: Baker RH (Ed.). Contaminants and Sediments. Ann Arbor
Press, Ann Arbor, Michigan, 522 pp.
Huntley, S.L., Wenning, R.J., Su, S.H., Bonnevie, N.L., Paustenbach, D.J. 1995.
Geochronology and sedimentology of the lower Passaic River, New Jersey. Estuaries
18(2):351-361.
Ingersoll, C.G., Dillon, T., Biddinger, R.G. 1997. Methodological uncertainty in sediment
ecological risk assessment. In: Ingersoll, C.G., Dillon, T., Biddinger, G.R. (Eds.).
Ecological Risk Assessments of Contaminated Sediment. SET AC Press, Pensacola,
Florida, 389 pp.
Long E.R., MacDonald D.D., Smith S.L., Calder F.D. 1995. Incidence of adverse
biological effects within ranges of chemical concentrations in marine and estuarine
sediments. Environmental Management 19:81-97.
Long, E.R., MacDonald, D.D. 1998. Recommended uses of empirically derived,
sediment quality guidelines for marine and estuarine ecosystems. J. Human Ecol. Risk
Assess. 4:1019-1039.
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O'Connor, T.P., Paul, J.F. 2000. Misfit between sediment toxicity and chemistry. Mar.
Poll. Bull. 40:59-64.
Pruell, R.J., Rubinstein, N.I., Taplin, B.K., LiVolsi, J.A., Norwood, C.B. 1990. 2,3,7,8-
TCDD, 2,3,7,8-TCDF, and PCBS in marine sediments and biota: Laboratory andfield
studies. USEPA ERL, Report to the USACE-NYD. Narragansett, RI. March 12.
Solomon, K.R., Ankley, G.T., Baudo, R., Burton, G.A., Ingersoll, C.G., Lick, W.,
Luoma, S.N., MacDonald, D.D., Reynoldson, T.B., Swartz, R.C., Warren-Hicks, W.
1997. Workgroup summary report on methodological uncertainty in conducting sediment
ecological risk assessments with contaminated sediments. Chapter 17. In: Ingersoll, C.G.,
Dillon, T., Biddinger, G.R. (Eds.). Ecological Risk Assessment of Contaminated
Sediments. SETAC Press, Pensacola, Florida. 389 pp.
U.S. Environmental Protection Agency (USEPA). 2001. The incidence and severity of
sediment contamination in surface waters of the United States: National Sediment
Quality Survey. Second Edition. EPA 823/F-01/031, Office of Water, Washington, D.C.
December.
U.S. Environmental Protection Agency (USEPA). 1997. Ecological Risk Assessment
Guidance for Superfund: Process for Designing and Conducting Ecological Risk
Assessments. Interim Final. EPA/540/R-97/006. Office of Solid Waste and Emergency
Response, Washington, D.C., June.
Wenning, R.J. 2001. Sediment Forum: USEPA forum on managing contaminated
sediments at hazardous waste sites: Summary of policy discussions. Contam. Soil Sed.
Groundwater, June/July, pp. 49-54.
Wenning, R.J., Dodge, D., Peck, B., Shearer, K. 1998. Screening-level ecological risk
assessment of PCDD/Fs and coplanar PCBs in lobster and sediments from Casco Bay,
Maine. Organo. Cmpds. 39:63-66.
Word, J.Q., Word, L.S. 2001. Hazardous contaminants in marine sediments. In: Lehr, J.,
Hyman, M., Gass, T.E., Seevers, W. (eds). Handbook of Complex Environmental
Remediation Problems. McGraw-Hill, NY. Chapter 5.
Section IV. A References:
Dickson, K.L., Maki, A.W., Brungs, W.A. 1987. Fate and Effects of Sediment-Bound
Chemicals in Aquatic Environments. Pergamon Press, New York, NY. pp. 449.
Ingersoll, C.G., Dillon, T., Biddinger, R.G. 1997. Methodological uncertainty in sediment
ecological risk assessment. In: Ingersoll, C.G., Dillon, T., Biddinger, G.R. (Eds.).
Ecological Risk Assessments of Contaminated Sediment. SETAC Press, Pensacola,
Florida, 389 pp.
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Paustenbach, D.J. 1995. Retrospective on U.S. health risk assessment: How others can
benefit. Risk Health Saf. Environ. (6):283-332.
Sharpe, S., Mackay, D. 2000. A framework for evaluating bioaccumulation in food webs.
Environ. Sci. Technol. 34(12):2373-2379.
U.S. Army Corps of Engineers (USACE). 1999. Ecological and Human Health Risk
Assessment Guidance for Aquatic Environments. Technical Report DOER-4. Dredging
Operations and Environmental Research Program, Waterways Experiment Station,
Vicksburg, MS. December.
U.S. Environmental Protection Agency (USEPA). 1999. Risk Assessment Guidance for
Superfund: Volume 3- (Part A, Process for Conducting Probabilistic Risk Assessment).
Draft. EPA/000/0-99/000. Office of Solid Waste and Emergency Response, Washington,
D.C. December.
U.S. Environmental Protection Agency (USEPA). 1998. Guidance for Submission of
Probabilistic Human Health Exposure Assessments to the Office of Pesticide Programs.
Draft. Office of Pesticide Programs, Washington, D.C., November 4.
U.S. Environmental Protection Agency (USEPA). 1997a. Policy for Use of Probabilistic
Analysis in Risk Assessment at the U.S. Environmental Protection Agency. Office of
Research and Development, Washington, D.C. May 15.
U.S. Environmental Protection Agency (USEPA). 1997b. Guiding Principles for Monte
Carlo Analysis. EPA/630/R-97/001. Risk Assessment Forum, Washington, D.C. March.
U.S. Environmental Protection Agency (USEPA). 1989. Risk Assessment Guidance for
Superfund. Vol. 1, Human Health Evaluation Manual (Part A). EPA/540/1-89/002.
Office of Emergency and Remedial Response. Washington, D.C. December.
Section IV, B1 References:
Agency for Toxic Substances and Disease Registry (ATSDR). 1995. Toxicological
Profile for Polycyclic Aromatic Hydrocarbons. Atlanta, GA. August.
Buamrad, P., Budzinski, H., Garrigues, P. 1998. Polycyclic aromatic hydrocarbons in
sediments and mussels of the western Mediterranean Sea. Environ. Toxicol. Chem.
17:765-776.
Cassarett and Doull. 1991. Casarette and Doull's Toxicology: The Basic Science of
Poisons. Fourth Edition. Pergamon Press, NY.
Eisler, R. 2000. Handbook of Chemical Risk Assessment. Lewis Publishers, Boca Raton,
FL. Chapter 25.
Huntley, S.L., Bonnevie, N.L., Wenning, R.J. 1995. Polycyclic aromatic hydrocarbon and
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petroleum hydrocarbon contamination in sediments from the Newark Bay estuary, New
Jersey. Arch. Environ. Contain. Toxicol. 28:93-107.
Huntley, S.L., Bonnevie, N.L., Wenning, R.J., Bedbury, H. 1993. Distribution of
polycyclic aromatic hydrocarbon (PAHs) in three northern New Jersey waterways. Bull.
Environ. Contam. Toxicol. 51:865-872.
U.S. Environmental Protection Agency (USEPA). 2002. Integrated Risk Information
System (IRIS). Cincinnati, OH.
U.S. Environmental Protection Agency (USEPA). 2000. Region 9 Preliminary
Remediation Goals (PRGs) 2000. San Francisco, CA. November.
U.S. Environmental Protection Agency (USEPA). 1999. Methods for Sampling and
Analyzing Contaminants in Fish and Shellfish Tissue. Office of Water. July.
U.S. Environmental Protection Agency (USEPA). 1993. Provisional Guidance for
Quantitative Risk Assessment of Polycyclic Aromatic Hydrocarbons. Office of Research
and Development. Washington, D.C. July.
r
U.S. Environmental Protection Agency (USEPA). 1989. Risk Assessment Guidance for
Superfund. Vol. 1, Human Health Evaluation Manual (Part A). EPA/540/1-89/002.
Office of Emergency and Remedial Response. Washington, D.C. December.
U.S. Environmental Protection Agency (USEPA). 1980. Ambient Water Quality Criteria
for Polynuclear Aromatic Hydrocarbons. Office of Water, Washington, D.C. EPA/440/5-
80/069.
Section IV, B2 References:
Abalos, M., Bayona, J., Compano, R., Granados, M., Leal, C., and Prat, M. 1997.
Analytical procedures for the determination of organotin compounds in sediment and
biota: critical review. Journal of Chromatography A, 788 (1-49)
California Environmental Protection Agency (Cal/EPA). 1999. Calculation of an
action/cleanup level for dibutyltin (DBT) at sites contaminated with tributyltin (TBT).
Department of Toxic Substance Control, Human and Ecological Risk Division.
December.
Environment Canada (EC). 1993. Priority Substances List Assessment Report. Non-
Pesticidal Organotin Compounds. Cat. No. En 40-215/18E. Ottawa, Canada, pp 32.
Extension Toxicology Network (ETN). 1996. Pesticide Information Profiles: Tributyltin.
June.
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Harino, H., Fukushima, M., and Kawai, S. 2000. Accumulation of butyltin and
phenyltin compounds in various fish species. Arch. Environ. Contam. Toxicol. July,
39(1):13-9.
Krone, C., Brown, A.D.W., Burrows, D.G., Bogar, R.G., Chan, S-L., Varanasi, U. 1989.
A method for analysis and measurement of butyltins in sediment and English sole livers
from Puget Sound. Mar. Environ. Res. 27:1-18.
Yamamoto, I. 1994. Pollution of fish and shellfish with organotin compounds and
estimation of daily intakes. Hokkaido Igaku Zasshi, March, 69 (2):273-81.
United States Environmental Protection Agency (USEPA) 1997. EPA holds National
Conference on Sediment Bioaccumulation in Contaminated Sediments News, number 18.
EPA-823-N-97-001.
United States Army Corps of Engineers (US ACE) 1998. Dredged Material Evaluation
Framework. November.
United States Environmental Protection Agency (USEPA). 1999. National Recommended
Water Quality Criteria. Office of Water. April.
Section IV. B3 References;
Canadian Council of Ministers of the Environment. 2001. Canadian sediment quality
guidelines for the protection of aquatic life. Polychlorinated dioxins and furans
(PCDD/Fs). In: Canadian Environmental Quality Guidelines. 1999. Winnipeg.
U.S. Environmental Protection Agency (USEPA). 2000a. Guidance for Assessing
Chemical Contaminant Data for Use in Fish Advisories, Volume 2: Risk Assessment and
Fish Consumption Limits, Third Edition. EPA 823/B-00/008. Office of Water.
Washington, D.C. November.
U.S. Environmental Protection Agency (USEPA). 2000b. Exposure and Human Health
Assessment of 2,3,7,8-Tetrachlorodibenzo-p-dioixn (TCDD) and Related Compounds.
Part II. Office of Research and Development, Washington, D.C. EPA/600/P-00/001Be.
September.
Van den Berg, M., Birnbaum, L., Bosveld, B.T.C., Brunstrorn, B., Cook, P., Feeley, M.,
Giesey, J.P., Hanberg, A., Hasegawa, R., Kennedy, S.W., Kubiak, T., Larsen, J.C., van
Leeuwen, F.X.R., Liem, A.K.D., Nolt, C., Peterson, R.E., Poellinger, L., Safe, S.,
Schrenk, D., Tillit, D., Tysklind, M., Younes, M., Waern, F., Zacharewski, T. 1998.
Toxic equivlanecy factors (TEFs) for PCBs, PCDDs, and PCDFs for humans and
wildlife. Environ. Health Perpsect. 106:775-792.
Section IV. C References:
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Coates, J.T., Elzermann, A.W. 1986. Desorption kinetics for selected PCB congeners
from river sediments. J. Contain. Hydrol. 1:191-210.
Eisler, R. 2000. Handbook of Chemical Risk Assessment. Lewis Publishers, Boca Raton,
FL. Chapter 25.
Jarman, W.M., Johnson, G.W., Bacon, C.E., Davis, J.A., Risebrough, R.W., Ramer, R.
1997. Levels and patterns of polychlorinated biphenyls in water collected from the San
Francisco Bay and estuary, 1993-95. Fresenius J. Anal. Chem. 359:254-260.
Lick, W, Rapaka, V. 1996. A quantitative analysis of the dynamics of the sorption of
hydrophobic organic chemicals to suspended sediments. Environ. Toxicol. Chem.
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Johnson, L. 2000. An Analysis in Support of Sediment Quality Thresholds for Polycyclic
Aromatic Hydrocarbons (PAHs)to Protect Estuarine Fish. Northwest Fisheries Science
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