Urittd!
Env«onmenm Protection
Agency
Office of Policy. Planning
and Evatoatan
EPA/230-10-88-041
November 1988
Review of
Ecological Risk Assessment
Methods
Exposure
Hazard
Risk Characterization
Receptor
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REVIEW OF ECOLOGICAL RISK ASSESSMENT
METHODS
Prepared By:
Sue Norton, Margaret McVey, Joanne Colt
Judi Durda, Robert Hegner
IGF Incorporated
9300 Lee Highway
Fairfax, Va. 22031-1207
Prepared For:
Dexter Hinckley
Office of Policy Planning and Evaluation
U.S. Environmental Protection Agency
401 M St. SW
Washington, D.C. 20460
November 1988
U.S. Environmental Protection Agency
Region 5, Library (PL-12J)
77 West Jackson Boulevard, 12th floor
Chicago, !L 60604-3590
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DISCLAIMER
This is a contractor's final report, which has been peer reviewed by the EPA
Offices and other contractors as described in the Acknowledgments. The
contents of this document do not necessarily reflect the views and policies of
•the U.S. Environmental Protection Agency, nor does mention of trade names or:'
commercial products constitute endorsement or recommendation for use.
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ACKNOWLEDGMENTS
Special acknowledgments are due to the reviewers of the draft report
who provided valuable comments and suggestions as well as specific
recommendations for the future use of ecological risk assessment in
regulatory agencies. In particular, we would like to thank Dr. Donald
Rodier of the EPA Office of Pesticides and Toxic Substances,
Environmental Effects Branch, Washington, D.C.; Dr. Ossi Meyn of the EPA
Office of Solid Waste and Emergency Response, Health Assessment Section,
Washington, D.C.; Dr. J. Garner of the'EPA Environmental Criteria and
Assessment Office in Research Triangle Park, North Carolina; Dr. Thomas
Barnwell, Jr., of the EPA Office of Research and Development (ORD),
Environmental Research Laboratory, in Athens, Georgia; Dr. Kenneth Perez
and Dr. Allan Beck of the EPA/ORD -Environmental Research Laboratory in
Narragansett, Rhode Island; Dr. Gerald Niemi of the EPA/ORD
Environmental Research Laboratory in Duluth, Minnesota; Dr. Steven
Lutkenhoff and Mr. Randall Bruins of the EPA/ORD Environmental Criteria
and Assessment Office in Cincinnati, Ohio; Dr. Denise Steurer of the EPA
Region V Water Quality Branch; Dr. Lawrence Barnthouse of the Oak Ridge
National Laboratory; Dr. Thomas Hallam of the University of Tennessee at'
Knoxville; Dr. John Conroy of COMECO; Mr. Dennis Logan of Marine
Ecological Research, Inc.; Ms. Carolyn Fordham of Environmental Science
and Engineering, Inc.; Dr. Patrick Sheehan of Aqua Terra Technologies;
and Dr. William Lappenbusch of Lappenbusch Environmental Health,
Incorporated.
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TABLE OF CONTENTS
Section Page
EXECUTIVE SUMMARY 1
1.0 INTRODUCTION 4
1.1 Purpose of the Report 4
1.2 General Approach and Scope 4
1. 3 Definitions 10
1.4 Organization of the Report 10
2.0 DISCUSSION AND CHARACTERIZATION OF METHODS 12
2 .1 Obj ective of Methods . 12
i:
2.2 Types of Methods and Applicability to Objectives . 13
2.2.1 Qualitative vs. Quantitative Methods 14
2.2.2 "Top-down" vs. "Bottom-up" Methods 19
2.3 Technical Characteristics 21
2.3.1 Definition of Receptors and Endpoints 21
2.3.2 Degree of Integration of Information
Concerning Multiple Chemicals
and Pathways 25
2.3.3 Treatment of Uncertainty 26
3.0 CONCLUSIONS AND RECOMMENDATIONS 31
3.1 Conclusions 31
3. 2 Recommendations 32
GLOSSARY 35
REFERENCES 40
APPENDIX A: Framework for Review of Legal Mandates
and Ecological Assessment Methods
APPENDIX B: Legal Mandates for Ecological Assessment
APPENDIX C: Ecological Assessment Method Summaries
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LIST OF TABLES
Number Page
1. Statutes, Directives, and Associated Methods
Reviewed 6
2. Additional Methods Reviewed 9
3. Guide to Methods Reviewed in Appendix C 11
4. Possible Endpoints for Ecological
Risk Assessment 22
LIST OF FIGURES
Number Page
1. An Integrated Model of Ecological Risk
Assessment 5
2. Quotient Method 15
3. Dose-Response Method 16
4. Extrapolation from Laboratory to Field 23
5. Uncertainty and Data Extrapolation 26
111
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REVIEW OF ECOLOGICAL RISK ASSESSMENT METHODS
KXJKUT.LVE SUMMARY
This document reviews several existing methods for conducting
ecological assessments to support regulatory decision-making. Ecological
risk assessment, strictly defined, is a procedure that predicts the
probability of adverse effects to ecosystems, or parts of ecosystems, from
environmental pollution. However, the term is often used to encompass many
types of ecological assessment procedures being used by EPA to support
regulatory decision-making. The methods covered in our review include
those that meet the strict definition of ecological risk assessment, as
well other types of methods used by state and Federal agencies. Overall,
ecological assessment methods reviewed in this document have been developed
for three major purposes: priority-setting, developing standards or
guidelines, and as input to risk management decisions. These methods are
in a constant state of flux, as existing methods are refined, new methods
are developed, and the data base needed to support the methods continues to
grow. :"
The purpose of this document is to identify general trends and
limitations of ecological assessment as it is currently practiced. We
reviewed 20 ecological assessment methods that have been used or are ready
to be implemented, and that we consider representative of ecological
assessment methods developed by EPA and other Federal and state agencies.
We have focused on methods designed to predict the likelihood and magnitude
of adverse effects that can result from releases of toxic or hazardous
substances into the environment, although several techniques for the
assessment of ecological damage are also included. Each method was
reviewed using a framework that we developed to ensure consistency among
the reviews. The framework consists of a set of characterization points in
each of the four major components of the ecological risk assessment
process: receptor characterization, hazard assessment, exposure assessment,
and risk characterization. We also reviewed the legislative mandates under
which the methods were developed, as well as the resources (data, cost,
level of expertise) needed to implement the methods.
Most of the methods included in our review are quantitative in nature,
although several were qualitative. The qualitative approaches can be used
to evaluate problems for which little quantitative information is
available, and can be effective for setting priorities, either as ranking
or screening procedures. They are generally less resource-intensive than
quantitative methods. Nevertheless, qualitative approaches are generally
not suitable for developing standards or managing risks; for these
applications, a quantitative approach is superior.
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The quantitative methods included in our review can be categorized into
quotient methods (those that provide a "yes/no" estimate of risk) and
continuous methods (those that provide an exposure-dependent estimate of
risk). Quotient methods are generally easily implemented, rely on readily
available data, and are well-suited for priority-setting and developing
standards. Risk management decisions generally require the more complex
and realistic continuous methods that provide an exposure-dependent
probabilistic estimate of risk. The continuous methods can be used for
priority- and standard-setting as well. The most significant drawback to
the use of continuous methods is that exposure-.response data are not
available for many pollutant/receptor combinations.
We found that the ecological assessment methods vary substantially with
respect to the receptors and endpoints evaluated. Most of the methods
address risks to populations rather than to communities or ecosystems, and
the endpoint of concern is often death, although reduced growth and
reproduction are also considered. This is primarily because data are often
insufficient to support risk assessment at the community and ecosystem
levels. For example, exposure-response data using community-level
endpoints such as species diversity are extremely limited. Exposure-
response data on population-level responses can be used to predict effects
at higher levels (i.e., a "bottom up" approach), but there is considerable
uncertainty associated with these types of predictions.
There are also differences among the methods in terms of their
treatment of uncertainty. Methods that rely on professional judgment do
not explicitly address uncertainty. A common way to treat uncertainty is
through the use of safety or assessment factors based on data adequacy;
this approach is often applied in methods used to set priorities or to
develop standards. Most of the risk assessment methods that incorporate
exposure-response information attempt to address uncertainty in both uhe
exposure and risk estimates. The techniques used include statistical
confidence limits, Monte Carlo simulations, sensitivity analysis, and field
validation and calibration.
We found that there are many trade-offs inherent to the ecological
assessment process. For example, the more realistic the method, the less
likely it is that adequate exposure-response data are available, and the
more difficult it becomes to propagate sources of uncertainty throughout
the analysis. Currently, the most realistic models used for probabilistic
risk assessment require inputs for biological parameters for which few
supportive data exist. Although the more complex analyses may more closely
approximate real-life conditions, their application may be limited by the
lack of supporting data and high operational resource requirements. On the
other hand, users of the more simple methods might not be aware of the
underlying assumptions that tend to limit the applicability of these
approaches.
Based on our review of the 20 ecological assessment methods, ICF and
the reviewers of our draft repc :t have several recommendations:
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Government-sponsored ecological damage assessments (field
investigations conducted to assess existing ecological damage)
provide an excellent opportunity for collecting data on the
relationship between pollutant levels and ecological damage. Agency
program offices should coordinate their efforts in this regard to
take full advantage of this data source.
Additional laboratory and field research is needed in several areas:
the interactions between stressors; sublethal effects; uncertainties
associated with extrapolation of laboratory data to the field; and
modeling of population responses to pollutants considering density-
dependent population interactions.
More emphasis should be placed on exposure-response evaluation, with
less emphasis on identifying "no effect levels" based on hypothesis
testing. In addition, more emphasis should be placed on statistical
treatment of uncertainty.
Endpoints should be chosen based on ecological, societal, and
regulatory significance as well as on data availability. Although
endpoints such as species diversity, altered nutrient cycling, and
ecosystem resilience may be more significant ecologically than
endpoints such as mortality, theoretical constraints and the lack of
exposure-response data currently limit the applicability of :'
ecosystem-level endpoints to most problems.
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1.0 INTRODUCTION
1.1 PURPOSE OF THE REPORT
This report reviews several existing methods developed by EPA and other
Federal and state agencies for assessing ecological impacts or risks
associated with the release of toxic or hazardous substances into the
environment. The purpose of this report is to identify general trends and
limitations of ecological assessment methods as they are currently
employed. This document is not intended to provide a step-by-step approach
for conducting ecological assessments, but to provide a baseline
understanding of the processes and general applications of ecological
assessment methodologies.
1.2 GENERAL APPROACH AND SCOPE
We reviewed several existing ecological assessment methods to identify
technical approaches to ecological assessment and to place these approaches
in the context of solutions to- particular needs, such as developing
standards or establishing testing priorities. To ensure a consistent
approach to each review, we developed a framework for the description of . •.
each method. Our reviews were structured to address the four basic
components of ecological risk assessment illustrated in Figure 1: (1)
receptor characterization, (2) hazard assessment, (3) exposure assessment,
and (4) risk characterization. We also described the operational resources
(e.g., cost, level of expertise) required to implement each approach.
Our review is limited to methods that we consider representative of
those in use by state or Federal agencies, or methods that are ready to be
implemented; we do not include those that are still experimental. We focus
on methods designed to predict the possibility or probability of adverse
ecological impacts. Methods designed to quantify existing damages or those
used to model exposure are included only in terms of their contribution to
predicting adverse ecological effects.
To understand the methods in the context of their intended application,
we also review the legislative or executive directives, under which the
methods were developed. Table 1 shows the legislative mandates considered,
the methods developed under each, and a brief summary of the purpose of
each method. Some of the methods covered in this document were not
developed under any specific legislative mandate, but are included because
they were developed for EPA use or had been used by EPA offices; these are
presented in Table 2. For simplicity, we refer to methods in this report
by the government office for which the method was prepared. The exceptions
to this are some of the methods which were not developed under
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Hazard
Exposure
..How much it released?
For how long?
How tar does it go?
T«n.fo,ma«on
r
L
Where does it end up?
How toxic is it
to different
species?
Life stages?
How does length
of exposure
affect toxicily?
How do other
factors affect
loxicily?
Toiicity
J
Chronic/Acute
Modifying
factors
tow much for
how long?
' Dose *" R««P
onse
Habitat
£ Whi
Trophic Level
where do they live? Risk Characterization
Life Stagu
How are eggs, juveniles, adults different?
Species I What are the communities ol concern?
Receptor
FIGURE 1. AM IMTBGRATED MODEL FOK ECOLOGICAL RISK ASSESSMENT
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HEIHOO NO
SIAIUIE 0« DIRECTIVE
TABLE 1: SIAIUIES. DIRECTIVES. AND ASSOCIATED METHODS REVIEWED
MANDATE METHOD
Legislation Administered by EPA
1- Comprehensive Environmental
Response, Compensation, and
Liability Act (CERCIA.
including the 1986 amendments)
2
Clean Air Act
to establish a National program for
responding to releases of hazardous
substances into the environment.
To protect and enhance the quality
of the Nation's air resources through the
prevention and control of air pollution.
BRIEF DESCRIPTION OF OBJECTIVE
Measuring Damages to Coastal and Marine
Environment Natural Resources (DOI
1987a>
To establish procedures for assessing
damages to natural resources from a
relatively small discharge of oil or a
release of a hazardous substance.
To establish procedures for assessing
damages to natural resources from a
relatively large discharge of oil or a
release of a hazardous substance.
Review of the National Ambient Air To assess the risks of ozone to human
Quality Standards for Ozone (EPA/OAR 1986) welfare.
Natural Resource Damage Assessments.
Final Rule (for Type I Assessments)
(DOI 1987b)
An Assessment of the Risk of Strato-
spheric Modification (EPA/OAR 1987)
To assess the ecological impacts of
chemicals that modify the stratosphere.
Clean UJtcr Act (including
the Water Quality Act of 1987)
To restore and maintain the chemical,
physical, and biological integrity of
the Nation's waters.
9-
10-
11
federal Insecticide,
Fungicide, and Rodenticide
Act (FIFRA)
lo prevent "unreasonable adverse effects
on the environment" from the misuse of
pesticides.
Guidelines for Deriving Numerical
National Water Quality Criteria for the
Protection of Aquatic .Organisms and their
Uses (EPA/OURS 1986) .
An Approach to Assessing Exposure to
and Risk of Environmental Pollutants
(EPA/OURS 1983)
Water Quality-Based Permitting for Toxic
Pollutants (EPA/OURS 198S, 1987)
Biological Criteria for the Protection
of Aquatic Life (Ohio EPA 1987a,b, 1988)
Nigara River Biota Contamination Project:
Fish Flesh Criteria for Piscivorous
Wildlife (NYS/OEC 1987)
Standard Evaluation Procedure for
Ecological Risk Assessment (EPA/OPP 1986)
Chemical Migration Risk Assessment
(Onishi et al. 1982, 1985)
To standanze a method for deriving
guidelines for chemical concentrations
consistent with the protection of
aquatic organisms, human health and
some recreational activities.
To provide guidelines for conducting
risk assessments for waterborne
pollutants.
To provide guidelines for issuing National
Pollutant Discharge Elimination System
(NPOES) permits.
To standardize collection and use of
biological measures of surface water
quality.
To develop limits for fish tissue residues
levels protective of fish-eating wildlife.
To outline procedures to assess risk of
pesticide uses proposed for registration.
To predict the occurrence and duration
of pesticide concentrations in surface
receiving agricultural runoff and
to predict the potential damage to aquatic
biota.
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TABU 1: STATUTES. DIRECTIVES. AND ASSOCIATED METHODS REVIEWED (continued)
METHOD NO.
STATUTE 00 DIRECTIVE
MANDATE
METHOD
BRIEF DESCRIPTION OF OBJECTIVE
12-
13
U
Marine Protection and
Resources Sanctuaries Act
Resource Conservation and
Recovery Act (RCRA)
To regulate the disposal of waste in the
ocean in order to protect human health.
welfare, and amenities, and the marine
environment, ecological systems, and
economic potentialities.
To promote the protection of health and
the environment and to conserve valuable
material and energy sources.
Safe Drinking Water Act
Toxic Substances Control Act
(ISCA)
To assure that all public drinking water
systems provide safe, high quality water.
To protect human health and the
environment through the regulation of the
manufacture and use of chemicals.
16-
We have not identified any methods
developed under this legislation.
Potential for Environmental Damage:
Proximity of Mine Waste Sites to Sensitive
Environments (EPA/OSW 1987a)
Technical Resource Document for
Risk-based Variances from the Secondary
Containment Requirement of Hazardous
Uaste Tank Systems. Volume II:
Risk-based Variance. (EPA/OSU 1987b)
The RCRA Risk-Cost Analysis Model, Phase
III Report (EPA/OSU 1964)
We have not identified any methods
developed under this legislation.
Estimating "Concern Levels" for Concen-
trations of Chemical Substances in the
Environment (EPA/OTS 1904)
Ecological Risk Assessment in The
Office of Toxic Substances: Problems and
Progress 1984-1987 (EPA/OTS 1987)
To aid in screening different mining
activities for potential environmental
impacts.
To determine environmental risks posed
by the release of uaste constituents
from hazardous waste tanks.
To aid in designing regulations
governing hazardous waste treatment
that are protective of human health and
the environment.
To aid in the evaluation of premanu-
factunng notices for new chemicals
which may or may not be toxic or
released into the environment.
To define potential hazards posed by
toxic chemicals.
Legislation Administered by Other agencies
Coastal Zone Management Act
of 1972
Endangered Species Act
Executive Order 11988,
Hoodplam Management
To encourage and assist States to develop
and implement mangement programs to
protect the land and water resources
of the coastal zone.
To conserve the ecosystems upon which
endangered or threatened species depend.
and to conserve the species themselves.
Requires that federal agencies take actions
to reduce the risk of flood loss, and to
restore and preserve the natural and
beneficial values served by floodplains.
We have not identified any methods
developed under this legislation.
We have not identified any methods
developed under this legislation.
We have not identified any methods
developed under this legislation.
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HE 1 HOD NO.
STATUTE Oft DIRECTIVE
I ABLE 1: SIAIUltS, DIRECTIVES. AND ASSOCIAUD METHODS REVIEWED (continued)
MANDATE METHOD
BRIEF DESCRIPTION Of OBJECTIVE
Executive Order 11990,
Protection of Wetlands
fish and Wildlife
Coordination Act
Requires that federal agencies take actions
to minimize the destruction, loss, or
degradation of wetlands, and to preserve
and enhance the natural and beneficial values
served by wetlands.
To assist in developing and protecting all
species of wildlife, wildlife resources, and
wildlife habitat.
National Environmental Policy Act lo preserve and enhance the environmental
(NEPA) quality of the Nation.
Wild and Scenic Rivers Act
To preserve selected rivers in their
free-flowing condition, to protect the
water quality of such rivers, and to
fulfill other vital national conser-
vation purposes.
We have not identified any Methods
developed under this legislation.
We have not identified any methods
developed under this legislation.
We have not identified any methods
developed under this legislation.
We have not identified any methods
developed under this legislation.
CO
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TABLE 2
ADDITIONAL METHODS REVIEWED
METHOD NO
METHOD
BRIEF DESCRIPTION OF OBJECTIVE
v£>
17- Methodology for Environmental Risk
Analysis (Barnthouse 1982); Users Manual
for Ecological Risk Assessment
(Barnthouse et al. 1986)
18- Regional Ecological Assessments: Concepts,
Procedures and Application (Ballou et al.
1981)
To support EPA/ORD's synfuels research
program by developing an environ-
mental risk assessment methodology.
To discuss the approaches developed
at Argorme National Laboratory to
conduct regional ecological assessments
for the impacts of alternative energy
opt i ons.
19- Computer Simulation Models of Assessment
of Toxic Substances (Eschenroeder et al.
1980)
20- Unfinished Business: A Comparative
Assessment of Environmental Problems
To develop a comprehensive chemical
fate and transport simulation model
which includes transport through the
biotic environment.
To estimate and rank current ecological
risks posed by 31 environmental problems.
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any specific legislation. More detailed information describing authors is
given, when available, in Appendix C and in the references. A quick
reference guide to Appendix C is provided in Table 3.
1.3 DEFINITIONS
Because ecological impact and risk assessment are still emerging
fields, a wide variety of terms and definitions have been used to describe
different types of analyses. In this document, we use the term "ecological
assessment" to refer to any type of assessment related to actual or
potential ecological effects resulting from human activities. The term
therefore encompasses impact, risk, and damage assessments as well as the
establishment of environmental criteria based on ecotoxicology.
The phrase "ecological risk assessment" is commonly used to cover this
array of analyses. In strict usage, however, ecological risk assessment is
a quantitative procedure that estimates the probability of specified levels
of ecological effects occurring in an ecosystem or part of an ecosystem in
response to a perturbation, and relates the magnitude of the impact to the
perturbation. In order to distinguish this more strict usage of the word
"risk," we refer to this latter type of analysis as "ecological risk
assessment" in this document and to all other analyses as "ecological a
assessments."
1.4 ORGANIZATION OF THE REPORT
The remainder of this report presents the major findings of our review.
In Section 2.0, we discuss the general purposes of ecological assessments
and the approaches currently used. We discuss our general conclusions and
recommendations for future developments in the area of ecological risk
assessment in Section 3.0. A glossary that includes terms used in the text
and appendices follows the text. Appendix A describes our approach to
reviewing the different methods and presents the framework used for each
review. Appendix B summarizes the requirements of 15 legislative or
executive directives for ecological assessments. Finally, Appendix C
presents the results of our detailed discussion of each method.
10
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TABLE 3
GUIDE TO METHODS REVIEWED
IN APPENDIX C
Method No. Method Pages
1 - CERCLA Type A Damage Assessment (DOI 1987a) C- 2- 9
2 - CERCLA Type B Damage Assessment (DOI 1987b) C-10-13
3 - Ozone Staff Paper (EPA/OAR 1986) C-14-16
4 - Stratospheric Modification (EPA/OAR 1987) C-17-21
5 - Ambient Water Quality Criteria (EPA/OWRS 1986) C-22-24
6 - Approach to Exposure and Risk ^EPA/OWRS 1983) C-25-27
7 - Water Quality-based Permitting (EPA/OWRS 1985, 1987) C-28-31
8 - Biological Criteria (Ohio EPA 1987a, 1987b, 1988) C-32-38
9 - Niagara River Fish Flesh Criteria (NYS/DEC 1987) C-39-41'
10 - Standard Evaluation Procedure for Ecological Risk
(EPA/OPP 1986) C-42-46
11 - Chemical Migration Risk Assessment
(Onishi et al. 1982, 1985) C-47-51
12 - Proximity to Sensitive Environments (EPA/OSW 1987a) C-52-53
13 - HWT Risk-based Variance (EPA/OSW 1987b) C-54-56
14 - RCRA - Risk Cost Analysis Model (EPA/OSW 1984) C-57-60
15 - Estimating Concern Levels (EPA/OTS 1984) C-61-62
16 - Ecorisk in OTS (EPA/OTS 1987) C-63-66
17 - Users' Manual for Eco. Risk (Barnthouse et al. 1982, 1986).. C-67-78
18 - Regional Ecological Assessments (Ballou et al. 1981) C-79-86
19 - Computer Simulation Model (Eschenroeder et al. 1980) C-87-88
20 - Comparative Risk Project (EPA/OPPE 1987) C-89-92
11
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2.0 DISCUSSION OF METHODS
In this section, we discuss the major findings of our review. In
Section 2.1, we identify several objectives of ecological assessments, and
provide examples of methods developed for each objective. In Section 2.2,
the methods are grouped broadly by the type of approach used. Finally, the
technical content of each method is discussed in Section 2.3.
2.1 OBJECTIVES OF METHODS
All of the methods included in our review were developed to assess the
potential for adverse effects on ecosystems or parts of ecosystems from the
release of hazardous substances into the environment. Within this overall
intent, however, the methods were developed for several practical
applications. The objectives of an ecological assessment method depend, in
part, on the legal mandate under which the method was developed. Most of
the legal mandates require consideration of "human health and the
environment". However, some mandates more specifically request the
development of standards or guidelines to limit concentrations of hazardous
substances in the environment (e.g., the Clean Water Act specifies the
"development of federal water quality criteria"). In other situations -.-
(e.g., energy development, pesticide use), adverse ecological effects must
be weighed against the benefits to human welfare. In these cases, a
quantitative ecological risk assessment may be used for comparing
incremental costs and benefits. Prioritization schemes have been developed
to aid decision makers in narrowing the large scope of regulatory concerns
and to target situations that require more testing or research.
The specific objectives of each method are summarized in Tables 1
and 2. In general, we found that most methods have one of three
objectives: (1) to set priorities, (2) to support the setting of standards
or guidelines, and (3) to assess risk as input to risk management
decisions. While some of the methods were developed specifically for one
of these objectives, several of them could be easily applied to more than
one objective.
Several of the methods included in our review are ranking schemes used
to set priorities. For example, the Comparative Risk Project (EPA/OPPE
1987) ranks broad environmental problems. The RCRA Risk-Cost Analysis
Model (EPA/OSW 1984) scores hazardous waste treatment, storage, and
disposal technologies with respect to their relative potential for
producing adverse ecological impacts. Other ecological assessment methods
have been used as screening procedures to identify priorities for further
testing or research. For example, the Standard Evaluation Procedure (SEP)
for Ecological Effects (EPA/OPP 1986) and the Method for Estimating Concern
Levels (EPA/OTS 1984) identify chemicals which, by their projected pattern
of release into the environment, are of sufficient ecological concern to
merit further testing. The Approach for Assessing Exposure and Risk
developed by EPA/OWRS (1983) describes a method of identifying combinations
12
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of plant and animal species, exposure levels, and projected effects on a
site-specific basis that are of sufficient ecological concern to merit
further investigation. Similarly, EPA/OSW (1987a) screened mine sites for
the potential to adversely affect valuable ecosystems based on to their
proximity to sensitive habitats. One well-known priority setting method
that we did not review is the Hazard Ranking System (MRS), which is used to
select sites for the Superfund National Priorities List. This is because
the current version of the system has a very limited ecological assessment
component, and the proposed revisions are not available for public comment
at the time of this publication.
Several of the methods reviewed were developed to assist in setting
standards or guidelines to limit releases of chemicals into the
environment. The Ambient Water Quality Criteria methodology has been used
to set guidelines for limiting ambient pollutant concentrations in surface
water to be protective of aquatic life. The Water Quality-based Permitting
for Toxic Pollutants (EPA/OWRS 1985, 1987) approach establishes guidelines
for using whole-effluent toxicity tests to establish effluent discharge
limits. In the Ozone Staff Paper (EPA/OAR 1986), EPA estimated
concentrations of ozone in the atmosphere that should be protective of
agricultural plants and forests. As an approach to calculating acceptable
release levels of chemicals, the EPA/OTS (1984) method to estimate Concern
Levels identifies concentrations of chemicals that may cause adverse . :
environmental effects in aquatic populations. The NYS/DEC (1987) used the
Niagara River Fish Flesh Criteria methodology to set limits for contaminant
residue levels in fish that would be protective of piscivorous animals such
as mink and osprey.
Most of the remaining methods were developed to conduct risk
assessments to support management decisions (e.g., Onishi et al. 1985,
EPA/OSW 1987b, EPA/OAR 1987). Of these, only the Land-Use Disturbance
portion of the Regional Ecological Assessments method (Ballou et al. 1981)
and the Stratospheric Ozone analysis (EPA/OAR 1987) evaluates risks
associated with agents other than toxic chemicals. The User's Manual for
Ecological Risk Assessment (Barnthouse et al. 1986) includes several
methods that can be used to support standard and priority setting as well
conducting quantitative risk assessments.
Three of the methods included in our review assess existing ecological
impacts. The Type B damage assessment (DOI 1987b) includes methods to
measure ecological damages that have already occurred. Although damage
assessment is not the focus of this report, these methods can provide
information that can be used as a baseline for predictive risk assessment,
and can offer insight into the appropriate receptors, endpoints, and
exposure pathways to be used in risk assessment. The Type A damage
assessment procedure (DOI 1987a) was developed to estimate damage, but can
be applied to predicting the occurrence of adverse effects. The Ohio EPA
(1987a, 1987b, 1988) Biological Criteria and evaluation methodology
includes measurements of fish and macroinvertebrate sub-communities that
reflect the biological integrity of aquatic ecosystems.
13
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Other methods, such as Regional Assessment Units (Ballou e_t al. 1981),
were developed to provide a common basis for discussion and are included in
our review only in the context of their contribution to risk assessment.
2.2 TYPES OF METHODS AND APPLICABILITY TO OBJECTIVES
In this section, we discuss the general types of methods reviewed, some
of the strengths and weaknesses of each, and their suitability for the
objectives discussed above.
2.2.1 Qualitative vs. Quantitative Methods
Although most of the methods reviewed are based on quantitative
analyses, several use qualitative analyses, as discussed below.
Qualitative Methods. One of the qualitative methods described in
Ballou eg al. (1981) was designed to-identify potential conflicts in
implementing certain energy resource development scenarios by delineating
endangered and threatened species habitat. The EPA/OSW (1987a) used a
similar approach in "Proximity of Mine Waste Sites to Sensitive
Environments". In this analysis, mine sites in different raining segments
were identified as posing more or less of a threat to highly valued
environments based on their proximity to endangered species habitats, "
wetlands, and National parks and forests.
Two of the qualitative methods, EPA/OPPE (1987) and EPA/OWRS (1983) use
professional judgment to evaluate ecological effects. The Comparative Risk
Project (EPA/OPPE 1987) used professional judgment to broadly rank a large
diversity of environmental problems. Similarly, in the OWRS approach to
exposure and risk (EPA/OWRS 1983), a great deal of professional judgment is
used to identify specific surface water locations that should be further
investigated because of a high potential for adverse ecological effects.
The strength of qualitative approaches is that chey can be used to
evaluate effects and problems for which there is little quantitative
information. By using professional judgment, many levels of information
can be integrated into the decision-making process. The qualitative
methods included in our review illustrate that these methods can be used
effectively to set priorities, either as screening or ranking procedures.
Qualitative methods, however, are limited in their use for either
developing standards or managing risks. The effort required to implement
qualitative methods can be low relative to quantitative computer-based
models. However, if the methods that rely on professional judgment are to
yield useful results, the assessors must be highly skilled.
Quantitative Methods. The quantitative methods reviewed can be
categorized as quotient or ratio methods (illustrated in Figure 2) and
continuous or exposure-response methods (illustrated in Figure 3).
14
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FIGURE 2
QUOTIENT METHOD
IF
- 1, regulate.
100%
Mortality
Response
(percent)
50
no effect
LC
60/AF
LC
90
Assessment
Factor
•Meet
Concentration
Figure 2. One use of the quotient method is to compare an estimated environmental
concentration (EEC) to a toxicity benchmark (e.g., LC^Q). Generally a few percent
mortality is considered acceptable, and the 1X50 is multiplied by an assessment factor (AF)
to estimate the level above which more than a few percent mortality would occur. If the
EEC/LCtjQ x AF equals or exceeds 1, then a decision to regulate the chemical or to
investigate the situation more tully is made.
-------
FIGURE 3
EXPOSURE-RESPONSE METHOD
100%
Mortality
Response
(percent)
Concentration
Figure 3. In the exposure-response method, estimated environmental concentrations are used
to predict the level of response (e.g., 40 percent mortality). Confidence limits around
the exposure-response curve define the uncertainty associated with the predicted level of
response.
-------
Quotient Methods:
Quotient methods are used to determine whether or not a specified level
of environmental contamination might be of ecological concern. Reference
concentrations intended to be protective of a given receptor are
established and are compared with estimated environmental concentrations
(EECs). For example, as illustrated in Figure 2, an "assessment factor"
(AF) can be applied to an acute LC$Q value* (solid vertical line, Figure 2)
to estimate a reference concentration that is approximately equivalent to a
LC^ value (dashed vertical line, Figure 2). Environmental concentrations
that exceed the reference concentration are considered to have potential
adverse effects.
Several of the methods that we reviewed use the quotient method,
including EPA/OPP's (1986) Standard Evaluation Procedure for Ecological
Risk Assessment, EPA/OTS's (1984) method for Estimating 'Concern Levels',
and EPA/OSW's (1987b) method of evaluating risk from hazardous waste tanks.
Quotient methods provide essentially a "yes or no" determination of
risk and are therefore well-suited for screening-level assessments. In
these cases, a decision concerning the level of risk that is considered
acceptable is required, but is contained within the derivation of a
reference criterion concentration. For example, the AWQC method (EPA/OWRS
1986) assumes that 1 percent adult mortality is an acceptable acute effect
and that 5 percent of the species present in an ecosystem can suffer more
than 1 percent adult mortality without the expectation of adverse ecosystem
effects.
The primary advantage of the quotient method is that it is a relatively
low cost, easily implemented method that often relies on data that are
readily available for many chemicals (e.g., *LC$Q, ID$Q values).
Consequently, quotient methods are well-suited for setting priorities as
well as for developing standards.
A limitation of the quotient method is that it does not predict the
degree of risk or magnitude of effects associated with specified levels of
contamination. This is a minor limitation when setting standards or
priorities. Some of the methods, however, attempt to address this problem;
these can be called "modified quotient methods." In Barnthouse et al.'s
(1986) Quotient Method and in EPA/OSW's (1987) Risk-based Variance
procedure, the conclusions are expressed as "no concern", "possible
concern", and "high concern", depending on whether the ratio of the
*• An LC5Q value is the concentration of a substance in water that is
associated with the death of 50% of the organisms in a laboratory bioassay.
This notation is also used to describe other proportions; for instance,
refers to the concentration estimated to kill 1 percent of the test
organisms. Similarly, an LD^Q value is the dose (usually expressed as
mg/kg body weight) that is estimated to kill 50 percent of the test
organisms.
17
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estimated environmental concentration (EEC) to the reference concentration
is <0.1, 0.1 to <10, or >10, respectively. Another modification of the
quotient method was used by Onishi et aj.. (1985); in this method, EECs were
compared to Maximum Acceptable Toxicant Concentrations (MATCs)^ and LC5QS.
If the EEC was below the MATC, the fish population was considered "safe".
If the EEC was above the LC5Q, the fish population was believed to be at
risk of increased mortality. If the EEC was between the MATC and the LC5o,
the fish population was considered to be at risk of sublethal effects.
Exposure-Response methods:
Exposure-response or continuous approaches are used to estimate the
magnitude of effect associated with an estimated exposure concentration.
The full exposure-response curve is used to estimate risk. For example,
one could use an exposure-response curve (illustrated in Figure 3) to
estimate the exposure concentration expected to produce 10, 20, 50, or 100
percent reduction of a trout population due to direct effects of a
contaminant.
Derivation of explicit risk estimates for a given endpoint of concern
requires, at a minimum, relating estimated environmental concentrations to
exposure-response information. An example of the use of exposure-response
information to estimate population-level effects is Ballou et al.'s (1981)
assessment of the effects on crop productivity of atmospheric sulfur ''
dioxide released from fossil fuel burning plants. In this approach, the
concentration of S02 at varying distances from the emission source was
modeled and compared with dose-response data for SC^-induced reductions in
crop yields. The estimated losses in crop productivity were then mapped
spatially to provide an estimate of the areal extent and degree of crop
reduction with increasing distance from the source of air emissions.
The exposure-response approach is well-suited for situations in which
an estimate of the magnitude of risk estimated to occur is needed to
support the setting of a standard, or in cases where a risk management
analysis will be conducted. For example, the EPA/OAR (1986) used an
exposure-response approach to aid in setting standards. Exposure-response
data were used to estimate crop yield reductions with increasing ambient
ozone concentrations. EPA/OAR selected a 10 percent reduction in yield as
the level of concern. The selection of a particular level of effect
represents a risk management decision. The exposure-response approach
allows a decision-maker to choose a level of concern based on economic and
human welfare considerations if necessary. In this way, the exposure-
response approach may be more useful for providing input to risk management
decisions than the quotient method.
•
* The MATC is the toxic chemical threshold concentration lying in a
range bounded at the lower end by the highest concentration having no
effect (NOEL) and at the higher end by the lowest test concentration having
a significant toxic effect (LOEL) in a life cycle or partial life cycle test
18
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The exposure-response approach has also been used, in the case of the
RCRA Risk-Cost Analysis Model (EPA/OSW 1984), to aid in setting priorities.
Thus, one of the strengths of this approach is that it can be applied to
many purposes. A limitation of the approach, however, is that exposure -
response data are not available for many chemical/receptor combinations.
2.2.2 "Top-Down" vs. "Bottom-Up" Methods
Most of the dose-response or quotient methods in our review addressed
effects to individuals or populations. There are two ways to approach the
evaluation of effects at the community and ecosystem level. The first
approach, called the "top-down" approach, directly assesses changes in
function and structure of communities and ecosystems (e.g., alterations in
species diversity or primarily production- rates). The second approach, the
"bottom-up" approach, uses laboratory data on effects at lower levels of
organization (e.g., mortality of individuals) to model changes at the
community or ecosystem level.
Top-down Methods. Several of the methods reviewed could be called top-
down methods. These include EPA/OPPE's (1987) Comparative Risk Project,
the RCRA Risk-Cost Analysis Method (EPA/OSW 1984), EPA/OSWs (1987a)
Sensitive Environment analysis, Ballou e_£,al.'s (1981) Land-use Disturbance
and Endangered and Threatened Species, and Ohio EPA's (1987a, 1987b, 1988)..
Biological Criteria methods. Most of these methods were used to aid in
setting priorities, however, and none included an ecosystem-level exposure-
response analysis.
One of the strengths of top-down approaches is that they evaluate
changes in communities and ecosystems directly. These changes, as opposed
to such effects as changes in the number of individuals in the population
of a single species, are easily defended by decision-makers as being
important. However, at the present time, few data exist to directly
predict the effects of chemicals on ecosystem-level properties such as
structure and function. More fundamentally, there is no accepted
definition of "ecosystem health," in part because ecosystems are dynamic,
changing with environmental conditions as organisms adapt or adjust to
stress (EPA/OPPE 1987). Thus, most of the predictive top-down approaches
that we reviewed either address habitat alteration or are based on
professional opinion.
The Ohio EPA (1987a, 1987b, 1988) Biological Criteria and evaluation
methodology represents a top-down approach to evaluating existing
ecological impacts. The criteria are biological indices based on several
measures of fish and macroinvertebrate community structure (e.g., species
diversity, species number) that reflect the biological integrity of aquatic
ecosystems. In this case, "ecosystem health" is defined as the value of
the indices measured for "least impacted11 surface water bodies in each of
several ecoregions in the state.
The RCRA Risk-Cost Analysis Model (EPA\OSW 1984) is an example of a
top-down predictive approach that attempts to be quantitative. In this
approach, a generic ecosystem exposure-response curve is constructed,
19
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spanning concentrations corresponding to minimal effects on sensitive
species at the lower end to ecosystem-level catastrophic effects at the
high end. The curve was based on four studies in which an ecosystem-level
exposure-response relationship could be constructed empirically. Hence,
there are many uncertainties associated with applying this curve to other
chemicals or situations.
Because of the lack of toxicological data available for directly
evaluating changes in communities or ecosystems, quantitative risk
assessment cannot currently be conducted in a "top-down" manner, with very
few exceptions. However, as illustrated by the methods we reviewed, top-
down approaches can be used effectively in setting priorities.
Bottom-up Methods. Three of the methods included in our review (DOI
1987a, Eschenroeder e_t ai. 1980, Barnthouse et ai. 1986) could be called
"bottom-up" methods. These methods estimate community-level effects using
computer models and laboratory data on the responses of individuals and
populations to chemicals. One of the strengths of bottom-up approaches is
that they provide insight into the processes that transfer effects between
different components of ecosystems (e.g., through a food web). However,
because these models require a great deal of site- and chemical-specific
data, the bottom-up approaches are most useful in supporting risk
management decisions on a site-specific basis. •'
The theoretical constraints of the modeling approaches can be daunting.
The ecological theory required to use system analysis to predict changes in
communities and ecosystems from more basic information is still in its
infancy, despite early recognition of its importance (Kickert and Miller
1979). Although the models that attempt to predict community responses
using individual and population-level laboratory data illustrate the
substantial progress that has been made in ecological risk assessment, they
may not incorporate (for practical or theoretical reasons) responses that
can greatly influence the conclusions drawn about risk. At least one of
the three methods that model biotic community-level effects considers each
of the following biotic community relationships:
• Transfer of energy and biomass from primary productivity through
several trophic levels and transfer of coxicant-induced reductions
in productivity at one trophic level through successive levels;
• Effects of reduced recruitment of juveniles to population age-
structure, productivity, and biomass;
• Transfer of toxicants through several trophic levels; and
• Changes in respiration, feeding, or grazing rates, susceptibility to
predation, mortality, emigration, and fecundity in species at one
trophic level affecting species in successive trophic levels.
20
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However, none of the three methods appear to consider other important
factors influencing community response:
• Density-dependent population regulation (e.g., density-dependent
changes in juvenile survivorship compensating for mortality caused
by pollution);
• Toxicant-induced changes in behavior (e.g., avoidance, disrupted
chemical communication among animals); and
• Incorporation of biotic and abiotic stressors (e.g., parasites and
extreme weather, respectively) and toxicant-induced changes in
responses to these stressors.
2.3 TECHNICAL CHARACTERISTICS
We reviewed each ecological assessment method with respect to five
broad factors: (1) receptor characterization, (2) hazard assessment, (3)
exposure assessment, (4) risk characterization, and (5) operational
resource requirements. Our detailed framework for reviewing the methods is
presented in Appendix A, and the results of our review of each method are
found in Appendix C. In this section, we discuss four major technical ...
characteristics that appear to be particularly important in determining the
scope and results of ecological assessment as it is currently practiced.
In general, the technical characteristics of ecological assessment methods
are similar for methods that were developed for similar objectives (e.g.,
setting standards) or that are based on the same general type of approach
(e.g., quotient method, exposure-response approach). Therefore, we discuss
the technical characteristics of the methods in the context of the overall
objectives for which the methods were designed.
2.3.1 Definition of Receptors and Endpoints
Receptors are the components of ecosystems that are or may be adversely
affected by a pollutant or other stress. Endpoints are the particular
types of impact or potential impact a chemical or other environmental
stress has on a receptor (e.g., death, decreased growth or productivity).
Because of the complexity of natural systems, it is difficult to assess
potential impacts to all receptors for all endpoints. Therefore,
ecological assessment methods select particular types of receptors and
endpoints to be "indicators" of potential harm to all components of the
system.
The range of potential indicators is enormous. Table 3 presents a list
of some potential endpoints for receptors at the individual, population,
community, and ecosystem level. Although there is no accepted definition
of ecosystem "health", some possible indicators of ecosystem resilience
(the ability of an ecosystem to recover from stress) include measurements
of species diversity, nutrient cycling, and productivity at the community
or ecosystem level (Levin et al. 1983, NRC 1981, EPA/OPPE 1987). Because
there is general consensus on adverse effects at the population level,
21
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(e.g., decreased reproduction or growth, increased mortality), the most
appropriate endpoints for use in risk assessment at this time may be at the
population level.
Few of the methods reviewed selected receptors and endpoints based
solely on ecological significance. In fact, in most cases the selection is
driven by practicality, as the receptors and endpoints selected tend to be
those for which the most toxicity data are available. For example, most of
the methods assess aquatic receptors. Part of the reason for this is that
there is generally less information on the exposure of terrestrial
organisms, or on the toxic effects of chemicals to terrestrial organisms.
The particular approaches to receptor and endpoint selection under the
various types of methods are discussed below.
Methods Used to Develop Standards or Guidelines. Of the ecological
assessment methods used to develop standards, the most common endpoints
selected were laboratory measurements of acute or chronic toxicity at the
species level. Most often the endpoint of concern for short duration
exposures is death. For longer duration exposures, the endpoints of
concern are ususally reduced growth, reproduction, and survival of
sensitive life- stages. Sublethal effects such as altered respiration and
behavior are seldom considered, although these types of impacts can have
substantial consequences for populations as well as for communities and
ecosystems. "
Typically, toxicity data for a sensitive effect (endpoint) and a
sensitive species (receptor) are selected, if possible, to represent the
ecosystem as a whole, although other factors such as the type, source, and
release pattern of the chemical, and potential pathways for fate and
transport through the environment, are also considered. The general
assumption is that if a most sensitive species is protected from direct
toxic effects of a chemical or effluent, the ecosystem will not be
adversely affected (EPA/OWRS 1986). It is often necessary, however, to use
species that can be tested reliably in the laboratory (EPA/OWRS 1985) .
The use of indicator species to represent the ecosystem is often the
approach taken in developing standards because it allows expeditious review
of a large number of chemicals, which often is required for chemical
regulatory programs. However, this approach is limited because endpoints
are selected based on the most sensitive effects in laboratory species;
sensitive effects in field populations may be different or may be altered
by density-dependent interactions in the field. Figure 4 describes some of
the factors that influence differences between field and laboratory
population responses to toxic chemicals .
Ohio EPA, however, has developed Biological Criteria for surface water
quality that are measures of fish and invertebrate community structure
(Ohio EPA 1987a, 1987b, 1988). The criteria are based on the values of the
community indices measured in reference environments selected as examples
of "least impacted" surface water bodies in each of several ecoregions in
the state. This approach avoids many of the sources of error associated
22
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TABLE 4
POSSIBLE KNDPOINTS FOR ECOLOGICAL RISK ASSESSMENT
INDIVIDUALS
Change in respiration
Change in behavior (e.g., avoidance of contaminated areas)
Inhibition or induction of enzymes
Increased susceptibility to pathogens
Decreased growth
Death (particularly important in the case of endangered
species, where the loss of even one individual is considered
significant)
POPULATIONS
Decreased genotypic and phenotypic diversity
Decreased biomass
Increased mortality rate
Decreased fecundity
Decreased recruitment of juveniles
Increased frequency of disease
Decreased yield
Decreased growth rates
COMMUNITIES
Decreased species diversity
Decreased food web diversity
Decreased productivity
Increased algal blooms
ECOSYSTEMS
Decreased diversity of communities
Altered nutrient cycling
Decreased resilience
23
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FIGURE 4
EXTRAPOLATION FROM LABORATORY TO FIELD -
DIFFERENCES BETWEEN LABORATORY STUDIES AND FIELD CONDITIONS
Parameter
Chemical
Characteristics
Concentrations
Organisms
Genetic variation
Interaction
Behavior
Laboratory Bioassay
• usually exposed to one chemical
• well-defined
• can be measured
• system is well-mixed
usually use selected genetic strain
usually one species is exposed
number of organisms is usually
controlled
organisms cannot avoid exposure
Field Exposure
• usually exposed to a mixture
• can be measured, but olten are
estimated
variable, system is
not well mixed
• high natural variability
• many species present
• density-dependent population
fluctuations
• organisms can avoid exposure,
emigration/immigration possible
Environment
Toxicity modifying
factors
parameters closely controlled
• parameters highly variable,
uncontrolled
-------
with extrapolating froa the laboratory into the field; however, it is not
always possible to identify the source of observed degradation of the
biological communities.
Methods Used to Set Priorities. Ecological assessment methods that are
used to set priorities also often use the indicator species approach
because the approach relies on readily available data (often LC5QS or
U>50s) and provides a common base (the indicator species and endpoint) by
which all chemicals can be compared. A few of the priority-setting methods
reviewed (EPA/OSW 1984, 1987a, EPA/OPPE 1987) explicitly characterized
receptors at a community or ecosystem level (e.g., wetlands, rivers,
forests). However, this approach is used less frequently because the data
base available to relate chemical stresses to changes in communities and
ecosystems is extremely limited and situation-specific. Therefore,
assessments of ecological impacts or potential impacts tend to be limited
to qualitative evaluations (EPA/OSW 1987a, EPA/OPPE 1987) or are based on
broad assumptions regarding the relationship between effects at the species
and ecosystem levels (EPA/OSW 1984)."
Methods Used to Assess Risk for Risk Management Decisions. Most of
the methods used for quantitative risk assessment characterize receptors
and endpoints at the species or'population level, in part because of the
more extensive data base on toxic effects at these levels. In an effort tot
identify ecologically important endpoints that might otherwise be
overlooked (e.g., avoidance of spawning areas), EPA/OTS (1987) includes a
fault tree analysis in their method.
A few of the methods used for risk assessment and risk management have
attempted to select receptors and endpoints at the community and ecosystem
level based on ecological significance. The NAAQS review of ozone effects
(EPA/OAR 1986) and the stratospheric modification assessment (EPA/OAR 1987)
discuss community and ecosystem-level effects qualitatively. Some of the
more recent computer models also have attempted to model effects at the
community and ecosystem level (Barnthouse eg al. 1986, Eschenroeder 1981,
DOI 1987a). The Barnthouse and Eschenroeder models rely in theory on
sublethal effects data, which, as mentioned previously, are often lacking.
In these cases, the lack of experimental data on sublethal effects is
circumvented by assuming a generic relationship between lethal and
sublethal effects.
2.3.2 Degree of Integration of Information Concerning Multiple Chemicals
and Pathways
Potential receptors can be exposed to multiple chemicals or other
stresses via multiple pathways (e.g., water, sediment, food, air). Current
approaches to integrating multiple pathways and chemicals when estimating
risk are related to the particular objective of the method.
Methods Used to Set Priorities. The methods used to set priorities
(EPA/OWRS 1983, EPA/OPPE 1987) that were based on professional judgment did
not explicitly address multiple chemicals or pathways, although these
issues presumably would be considered by the individuals conducting the
25
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analysis. The Standard Evaluation Procedure for Ecological Risk (EPA/OPP
1986) is a chemical-specific method, but considers a different dominant
route of exposure for different types of receptors (aquatic and
terrestrial). Similarly, because the EPA/OTS 1984 method for Estimating
Concern Levels was developed to screen chemicals for further testing, the
analysis is conducted for single chemicals only. The RCRA Risk-Cost
Analysis model (EPA/OSW 1984) was developed to rank risks associated with
managing multi-chemical waste streams. The toxicity of the waste stream in
this method is based on the most toxic constituent of the waste stream.
••
'- Methods Used to Develop Standards or Guidelines. Multiple pathways and
chemicals are generally not considered in methods for standards development
primarily because standards usually address a single chemical in a given
medium. In the AWQC methodology (EPA/OWRS 1986), exposure to predators
through the food chain is addressed in addition to exposure to aquatic
organisms through the water. However, the two pathways are not integrated
(i.e., predatory fish are not exposed to both food and water); the most
protective value is chosen as the water criterion. Similarly, the NYS/DEC
(1987) Fish Flesh Criteria methodology can be used to develop residue
limits for single chemicals only. The Water Quality-based Permitting for
Toxic Pollutants guidelines (EPA/OWRS 1987), however, recommend assessing
the toxicity of whole effluents using bioassays. The Biological Criteria
for fish and macroinvertebrate communities developed by Ohio EPA (1988), on
the other hand, reflect aquatic community responses to multiple chemicals •'•
and other physical stresses via all routes of exposure.
Methods Used to Assess Risk for Risk Management Decisions.
Consideration of multiple chemicals and pathways are particularly important
for risk assessments conducted to support risk management decisions at
hazardous waste sites. At these sites, biota can be exposed to many
different chemicals through several different environmental media. Only
one of the methods we reviewed, however, addresses the effects of multiple
chemicals (EPA/OSW 1987b). Under the hazard index approach, the ratios of
estimated environmental concentrations to toxicity criteria are summed to
provide an index for total risk. This approach assumes that toxicity is
. additive and, theoretically, is therefore appropriate for those chemicals
with similar modes of action. Although the Ecosystem Uncertainty Analysis
(Barnthouse et al. 1986) method does not model multiple chemicals, it does
attempt to incorporate the effects of stresses in addition to the toxicant.
None of methods reviewed examine multiple pathways, although the CERCLA
Type A Assessment (DOI 1987a) models a different dominant route of exposure
for different types of receptors.
2.3.3 Treatment of Uncertainty
' There are many sources of uncertainty associated with ecological
assessments. For example, each component of risk assessment (i.e.,
, receptor, toxicity, and exposure assessment) has some uncertainty
associated with it, and when these components are combined to estimate
risk, new layers of uncertainty are added. Figure 5 illustrates how
uncertainties associated with each component of ecological risk
26
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FIGURE 5
UNCERTAINTY AND DATA EXTRAPOLATION
Increasing Uncertainty about effects
N>
Known
Chemical
Similar
Chemical
Controlled i
Environment
t Natural
Environment
Surrogate
Test
Organism
Ecosystem
Level Effects
Community
Level Effects
Population
Level Effects
Other
Organism
Hazard
Extrapolation
Exposure
Extrapolation
Receptor
Extrapolation
-------
assessment can combine to increase the uncertainty of risk predictions. A
thorough discussion of sources and treatment of uncertainty is beyond the
scope of this document. In this section, we briefly discuss the way in
which the methods reviewed treat uncertainty. The discussions are
presented in the context of the purposes for which the various methods were
developed.
Methods Used to Set Priorities. Two of the methods reviewed rely on
professional judgment to set priorities (EPA/OPPE 1986, EPA/OWRS 1983).
These methods do not treat uncertainty explicitly. The other methods for
establishing priorities (EPA/OPP 1986, EPA/OTS 1984, EPA/OSW 1984) address
uncertainty through the use of "application" or "assessment" factors. The
assessment factors are used to derive an environmental concentration level
which, if equaled or exceeded, is likely to elicit an adverse ecological
effect. Assessment factors are based on the adequacy of available data;
chemicals with limited toxicity data are assigned more conservative
application factors. EPA/OTS (1984) has used the assessment factor
approach extensively and has expended, considerable effort in estimating
appropriate factors for different sources of uncertainty. In addition,
EPA/OTS (1984) uses quantitative structure activity relationships (QSAR),
when appropriate, to extrapolate from tested chemicals to untested
chemicals. The RCRA Risk-Cost Analysis project (EPA/OSW 1984) uses safety
factors to account for the quality of data used to determine the ecosystem
threshold concentration. A safety factor is similar to an application •••
factor, except that the safety factor is intended to provide an estimate of
the boundary between a no effect concentration and an effect concentration.
Methods Used to Develop Standards or Guidelines. A major source of
uncertainty for methods used to develop standards or guidelines is in the
definition of a reference toxicity value. There are many sources of
uncertainty associated with the use of experimentally-derived reference
toxicity values. Some of the largest sources of uncertainty are associated
with extrapolations of data from a studied chemical, species, or time frame
to unstudied chemicals, different species or communities or ecosystems, or
longer time frames. In addition, there are uncertainties associated with
the extrapolation of responses in laboratory species to responses of
receptors in the field. There are also uncertainties associated with the
statistical methods used to derive toxicity values.
In the methods we reviewed, the reference toxicity values used to
develop standards or guidelines were derived from experiments designed to
test a hypothesis (e.g., define a no-effect level, chronic values in
EPA/OWRS 1986) or derived from exposure-response data (EPA/OAR 1986, acute
values in EPA/OWRS 1986). Each of these approaches has uncertainties
associated with it. The size of the experimental error depends in part on
the total sample size used to derive the value. Because the total sample
size needed to construct an exposure-response curve is usually larger than
that used in hypothesis testing, the error should be smaller for a toxicity
value derived from exposure-response data. For this reason, when they are
available, reference toxicity values derived from exposure-response data
are preferable to values derived from hypothesis testing.
28
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Another common source of uncertainty is associated with deriving
reference toxicity values for chemicals that have not been tested
extensively. To avoid the uncertainty associated with taxonomic variation,
EPA/OWRS (1986) does not develop AWQC unless minimum data requirements for
taxonomic breadth are met. However, as in the methods used to set
priorities., the most common treatment of uncertainty is to apply generic
assessment or safety factors to a toxicity value. This approach is used to
extrapolate from acute data to chronic no- or lowest-observed-effect levels
(EPA/OWRS 1986), from studied species to unstudied ones, and from the
effects seen in the laboratory to those expected in the field (EPA/OTS
1984).
In the case of extrapolation from acute to chronic effects, Barnthouse
e_£ al. (1982, 1986) recommends quantifying uncertainty more rigorously,
through a statistical analysis of the extrapolation error. For example,
under this approach, uncertainty of the extrapolation between acute (LC$Q)
and chronic (MATC) toxicity data is determined by regression of the two
values for many species. The error estimate is then defined by the
variance associated with the MATC-LC50 regression.
Methods Used to Assess Risk for Risk Management Decisions. In the
methods that quantify risk, uncertainty arises not only in conjunction with
toxicity values, but also in parameters and models used to estimate -.-
exposure and in the method used to integrate exposure and toxicity
information into an estimate of risk.
Several techniques have been employed to address uncertainty in
exposure estimates. DOI's (1987a) CERCLA Type A Assessment uses only one
point (the mean) to represent the value of each environmental parameter.
EPA/OSW (1987b) uses two values to represent a parameter: the average and
the plausible worst case (e.g., upper 95th percentile). EPA/OAR (1986)
used statistical confidence limits to bound long-term ozone concentrations
based on confidence limits of the original (short-terra) field measurements
and the unexplained variation in the model regression used to predict long-
term concentrations.
For reasonably complex models, a strict analytic accounting of each
source of uncertainty becomes extremely labor-intensive. Therefore, more
complex models (e.g. Onishi e_£ al. 1982, 1985; Barnthouse, 1982, 1986) do
not attempt to define statistical confidence limits analytically but
address uncertainty using a variety of techniques. These techniques
include th« use of probability distributions as input variables (Onishi et.
al. 1982, 1985), Monte Carlo simulations (e.g., randomly combining
variables chosen from several distributions; Barnthouse 1982, 1986),
sensitivity analysis, calibration, and validation. Although a rigorous
discussion of these techniques is beyond the scope of this document, these
techniques, as used in the methods reviewed, are discussed briefly below.
The Monte Carlo simulation technique, used by Barnthouse e_t al. (1986),
is used to represent multiple sources of environmental uncertainty. A
probability distribution is measured or estimated for each of the parameter
values. For each run of the Monte Carlo simulation, a single value is
29
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selected at random for each parameter from a defined probability
distribution. This procedure is iterated until a reasonably stable
probability distribution (convergence) is obtained for the results. The
variance of the results can then be characterized. Monte Carlo simulation
is a particularly useful technique for combining multiple sources of
variance if parameters are not normally distributed and if the functional
form of the model is non-linear. However, Monte Carlo simulation can
require a long time to obtain convergence, perhaps days or months.
For important parameters that are not well-characterized (e.g.,
indirect effects of toxicants), sensitivity analyses can help define a
range of possibilities for the final predictions. For example,
Eschenroeder et al. (1981) analyzed the effect of variation in toxicant-
induced mortality rates of primary producers and feeding and respiration
rates at various other trophic levels on the predictions of biomass
reduction at each trophic level. A. sensitivity analysis has also been
performed on the surface water component of Onishi et ai.'s (1982, 1985)
Chemical Migration Risk Assessment methodology.
The techniques discussed above help to define the uncertainty
associated with a model's predictions. Other techniques help to reduce
model uncertainty by increasing accuracy and precision. Validation
consists of assessing the accuracy of the model by comparing model inputs
and predictions with field data. Once model accuracy is defined, "
calibration of the model or submodels with field data can improve model
accuracy. During calibration, parameter values are varied until a close
match is obtained between the model predictions and the field measurements.
The Ecosystem Uncertainty Analysis (Barnthouse et al. 1986) has been
partially calibrated using experimental ponds, although the model itself
was also modified to more closely resemble the dynamics of the ponds (as
described in EPA/OTS 1987). Several of the other models (e.g., EPA/OAR
1987, Eschenroeder et al. 1981) could be calibrated using field or
historical data. For example, the relationships between bioaccumulation of
chemicals and octanol/water partitioning coefficients used in Eschenroeder
et al. (1981) could be validated using measured chemical concentrations in
tissues of animals at various trophic levels.
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3.0 CONCLUSIONS AND RECOMMENDATIONS
3.1 CONCLUSIONS
Comparison of the 20 methods revealed that each method addresses one or
more of the components of probabilistic risk assessment and that the scope
and endpoints under consideration vary considerably. Some methods were
designed primarily to aid in the setting of priorities (EPA/OSW 1987a,
EPA/OPPE 1987), others to support the development of criteria and standards
(EPA/OWRS 1986, EPA/OAR 1986, EPA/OTS 1984), and most of the others to
provide input to risk management decisions (Eschenroeder e_£ al. 1980,
Barnthouse e_t al. 1986). Nonetheless, a few general conclusions can be
drawn from this review:
• Methods can be grouped as qualitative or quantitative approaches;
quantitative methods can be further grouped into quotient (ratio)
methods or exposure-response (continuous) methods.
While qualitative methods cannot be used to- develop standards
or to quantify risk, they can be very effective as screening o'r
ranking tools to set priorities.
The quotient method (yielding dichotomous predictions)
appropriately addresses legislative mandates to identify "safe"
levels of chemicals.
Exposure-response methods (yielding continuous predictions) may
be more useful for risk management decisions when it is
important to know the degree of damage anticipated with a
specified degree of exposure.
• Methods that address changes at the community or ecosystem level can
be grouped into "top-down" approaches or "bottom-up" approaches.
There is considerable uncertainty associated with inferring
community- and ecosystem-level effects from laboratory
bioassays (bottom-up approaches). Predictive tools for
combining density-dependent population regulation and density-
independent processes (e.g., responses to toxic chemicals) are
not yet available.
There is also uncertainty associated with addressing ecosystem-
level effects directly (top-down approaches). Exposure-
response data using community- or ecosystem-level endpoints
(e.g., species diversity and abundance) are extremely limited.
Because of the limited data, these methods have generally been
qualitative, and have been developed to help set priorities, or
to document existing impacts.
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Both types of approaches are limited in that there is no
accepted definition of "ecosystem health," although
characterization of appropriate reference environments might
serve as an appropriate surrogate definition.
• As models become more complex and representative of real-world
processes, other difficulties arise.
The more complex the model, the more difficult it becomes to
propagate sources of error throughout an analysis. Other
techniques must be applied including validation, calibration,
and sensitivity analyses.
The more realistic the model, the less likely it is that
adequate toxicological data are available. Currently, the most
realistic models require .estimates for biological parameters
for which few supportive data exist. In particular, these data
include sublethal responses to toxicants and the interactions
between stressors (e.g., toxicant and temperature tolerance).
3.2 RECOMMENDATIONS
Based on this review, we recommend the following:
• Government-sponsored ecological damage assessments and monitoring
efforts should be coordinated to provide usable data for ecological
exposure-response modeling and exposure assessment. Points for
coordination include:
Prescribing chemical and physical data collection (e.g.,
sampling protocol for contaminants in different media,
including biological samples);
Prescribing biological data collection (e.g., species
diversity, nutrient flux, indicator species, time course of
response and recovery);
Standardizing and coordinating compilation of damage
assessments;
Comparison of monitoring data with predictions of ecological
changes that might have been made prior to the initiation of
monitoring (i.e., to verify or to refute specific modeling
efforts based on subsequent observations);
Use of monitoring data to rigorously define "healthy"
ecosystems for different geographic regions of the country;
Sponsoring data analysis for extraction of general ecological
and toxicological principles.
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Further laboratory- and field-based research is needed on several
issues:
Interactions between stressors in animal laboratory toxicity
testing;
Toxicant effects on sublethal endpoints such as respiration,
feeding rates, susceptibility to disease, and decreased
tolerance of other stressors;
Defining the risks associated with multiple chemical stresses
or multichemical waste systems. The reliability of three
approaches to this problem could be compared: (1) selecting the
single most toxic constituent to represent the mixture, (2)
hazard index (EPA/OSW 1987b), and (3) bioassays on the
mixtures;
Developing models that combine both density-dependent
population interactions and density-independent processes
(e.g., responses to environmental contamination) in a useful
predictive tool.
Future development of ecological risk assessment methods should : *
address several needs:
Increased emphasis should be placed on statistical treatment of
uncertainty;
When practicable, emphasis should be placed on exposure-
response evaluation with less emphasis on identifying "no-
effect-levels" based on hypothesis testing;
Most ecological assessments deal primarily with direct acute
effects; increased emphasis on indirect and chronic effects is
needed, especially for non-point sources of contamination;
There is a need to identify indicators of ecosystem health and
to focus on ecosystem resilience and recovery in order to
define what constitutes an acceptable level of effect and to
focus ecosystem modeling efforts and environmental monitoring
efforts most effectively;
Data requirements continue to pose a problem for all but the
most simple methods. To improve the utility of models that
predict community-level and ecosystem-level effects, additional
emphasis should be placed on the role of microcosm studies in
supplying input data and in calibrating model outputs.
Endpoints should be chosen based on ecological, societal, and
regulatory significance, as well as practicality.
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There is a need for ecosystem-level modeling efforts and more
emphasis on quantitative probability-based ecological risk
assessment methodologies.
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GLOSSARY
The following definitions were compiled primarily from these sources: Rand
and Petrocelli (1985), Krebs (1978), Odum (1971), Whittaker (1975),
Webster's New Collegiate Dictionary, and the McGraw-Hill Dictionary of Life
Sciences.
Anadromous: Organisms that ascend rivers from the sea to spawn or to
breed.
Assessment factor: A factor that is used to derive an environmental
concentration level from a specified toxicity benchmark (e.g.,
LC50), which if equaled or exceeded is likely to cause an adverse
ecological effect.
Benthos: Organisms that live on or'in the bottom of bodies of water.
Bioaccumulation: The net uptake of chemicals by organisms directly from
water or through consumption of food containing the chemicals.
Bioconcentration: The net uptake of chemicals by aquatic organisms from. .
water.
Biomagnification: The net increase in chemicals in organisms at
successively higher trophic levels as a consequence of ingesting
contaminated organisms at lower trophic levels.
Calibration: The adjustment of model parameter values using field data.
Community: An assemblage of populations of plants, animals, bacteria,
and fungi that live in an environment and interact with one another,
forming a distinctive living system with its own composition,
structure, environmental relations, development, and function.
Copepod: Any of a large class of microscopic freshwater and marine
crustacea; also, one of the most abundant types of animal in marine
zooplankton.
Demersal: Living at or near the bottom of the sea.
Density dependence: A density-dependent effect alters the birth rate or
the death rate of a population as a function of the density of the
population. Examples of factors that can have density-dependent
effects include competition among members of the population, rates
of attack by parasites and predators, and emigration.
Deterministic: A unique output or result produced by a specified input
(stochastic or random selection of parameters from probability
distributions are not incorporated).
35
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Diel: A 24-hour period, usually including the day and following night.
EC50: Median effective concentration. The concentration of material in
water that is estimated to be effective in producing some sublethal
response in 50% of test organisms. Often used in reference to
immobilization of invertebrates or reduced growth of algae.
Ecological damage assessment: A subcategory of ecological impact
assessment that measures effects that have already occurred in an
ecosystem or any part of an ecosystem as the result of a
perturbation. As the phrase is used in this report, damage
assessments can measure injury (i.e., adverse effects that may not
have economic consequences) as well as damage (i.e., adverse effects
that can be measured in economic terms). Measurements of damages
that have already occurred can be used as input into probabilistic
ecological risk assessments.
Ecological impact assessment: A subcategory of environmental impact
assessment that addresses impacts on non-human biota by documenting
and/or estimating the occurrence of impacts within an ecosystem or
any part of an ecosystem, and the consequences of the occurrence.
Ecological risk assessment: A subcategory of ecological impact -.--
assessment that (a) predicts the probability of adverse effects
occurring in an ecosystem or any part of any ecosystem as'a result
of a perturbation and (b) relates the magnitude of the impact to the
perturbation.
Ecological risk management analysis: A decision-making process that
considers political, social, economic, and engineering information
in conjunction with risk-related information to select an
appropriate regulatory response to a potential ecological problem.
EEC: Estimated environmental concentration.
Environmental impact assessment: A broad field that includes all
activities that attempt to analyze and evaluate the effects of human
action on natural and anthropogenic environments (after Suter e_t al.
1987).
Ecosystem: The biotic community and its environment which, together,
function as a system of complementary relationships, with the
transfer and circulation of energy and matter.
Environmental impact assessment: A broad field that includes all
activities that attempt to analyze and evaluate the effects of human
action on natural and anthropogenic environments (after Suter et al.
1987).
Environmental Impact Statement (EIS): An assessment of the impacts of a
specific Federal action as required by the National Environmental
Policy Act (NEPA). Although an EIS may use ecological risk
36
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assessment as a tool for comparing alternatives, an EIS is required
only to provide a "full and fair discussion of significant
environmental impacts".
Environmental risk assessment: A procedure that predicts (a) the
possibility of adverse effects occurring to human health or the
environment and (b) the consequences of the adverse effects (after
NRG 1975).
Fecundity: Potential capability of an organism to produce viable
offspring.
GMATC: The point estimate of the Maximum Acceptable Toxicant
Concentration (MATC) calculated as the geometric mean of the no-
effect level (NOEL) and the low-effect level (LOEL) taken from life-
cycle or partial life-cycle toxicity tests.
Hardness: The concentration of all'divalent metallic cations, except
those of the alkali metal, present in water. In general, hardness
is a measure of the concentration of calcium and magnesium ions in
water and is frequently expressed as mg/1 calcium carbonate
equivalent.
Hazard assessment: A component of risk assessment that consists of the
review and evaluation of toxicological data to identify the nature
of the hazards associated with a chemical, and to quantify the
relationship between dose and response.
Ichthyoplankton: The drifting eggs and larvae of many species of fish.
LC50: Median lethal concentration. The concentration of substance in
water that is associated with the death of 50% of the organisms in a
laboratory bioassay. This notation is also used to describe other
proportions; for instance LC^gt ^25, and LC^ refer to the
concentrations estimated to kill 10%, 25% and 1% of the test
organisms, respectively.
LOEL: Lowest observed effect level. The lowest level or concentration
of a material used in a toxicity test that produces a statistically
significant adverse effect on the exposed population of test
organisms as compared with the controls.
MATC: The toxic chemical threshold concentration lying in a range
bounded at the lower end by the highest tested concentration having
no effect (NOEL) and at the higher end by the lowest tested
concentration having a significant toxic effect (LOEL) in a life
cycle (full chronic) or partial life cycle (partial chronic) test.
Microcosm: A laboratory simulation of a portion of an ecosystem (e.g., a
terrarium).
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Mesocosm: A composite physical and biological model of an ecosystem,
intermediate in scale between a microcosm and a macrocosm, with a
level of organization as high as the natural world (e.g., ponds).
Macrocosm: A composite physical and biological model of an ecosystem,
large in scale, with a level of organization as high as the natural
world (e.g., an experimental watershed including terrestrial and
aquatic ecosystems).
Macroinvertebrates: Invertebrate species that are sufficiently large to
be handled without the aid of a microscope.
Monte Carlo simulation: An iterative modeling technique where parameter
values are drawn at random from defined probability distributions, a
solution defined, and the process repeated until a stable
distribution of solutions results.
NOEL: No observed effect level. The highest concentration in a toxicity
test that has no statistically significant adverse effect on the
exposed population of test organisms as compared with the controls.
Niche: Role or "profession" of an organism in the environment; its ac-
tivities and relationships in the community. •• •.
Piscivorous: Fish-eating.
Phytoplankton: Drifting aquatic plants, usually uni-cellular.
Phytotoxic: Toxic to plants.
Planktivorous: Eats drifting aquatic organisms.
Plankton: Microscopic or small macroscopic plants and animals which live
freely in surface water and which, because of their limited powers
of locomotion, drift with the water currents.
Population: A potentially interbreeding group of individuals of a single
species.
Primary productivity: The rate at which radiant energy is stored by
photosynthetic and chemosynthetic activity of producer organisms
(chiefly green plants) in the form of organic substances.
QSAR: Quantitative structure activity relationship: a method of
estimating unmeasured physical and toxicological properties for a
chemical on the basis of chemical structure, functional groups, and
similarity to known chemicals.
Receptor: The entity (e.g., organism, population, community, ecosystem)
that might be adversely affected by contact with or exposure to a
substance of concern.
38
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Resilience: The ability of a system to recover from perturbations.
Resistance: The ability of a system to absorb an impact without
significant change from normal fluctuations.
Risk management: The process of weighing policy alternatives and
selecting the most appropriate regulatory action based on the
results of risk assessment.
Secondary productivity: Rate of energy storage by heterotrophic
organisms (i.e., rate of energy storage at consumer levels -
herbivores, carnivores, or detritus feeders).
Species: A group of closely related, morphologically similar individuals
which actually or potentially interbreed.
Stochastic: A process involving a random variable.
Stratosphere: The upper portion of the atmosphere, in which temperature
varies very little with changing altitude and clouds are rare.
Trophic level: Functional classification of organisms in a community
according to feeding relationships; the first trophic level includes,
green plants; the second trophic level includes herbivores; and so
on.
Validation: The testing of a model against reality.
Zooplankton: Drifting aquatic animals.
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Environmental Protection Agency/Office of Solid Waste (EPA/OSW). 1987a.
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Environmental Protection Agency/Office of Solid Waste (EPA/OSW). 1987b.
Technical Resource Document for Variances from the Secondary
Containment Requirement of Hazardous Waste Tank Systems. Volume II:
Risk-based Variance. Environmental Protection Agency, Washington, D.C.
Environmental Protection Agency/Office of Toxic Substances (EPA/OTS). 1984.
Estimating "Concern Levels" for Concentrations of Chemical Substances
in the Environment. Environmental Effects Branch, Health and
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Environmental Protection Agency/Office of Water Regulations and Standards1'
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Krebs, C.J. 1978. Ecology: the Experimental Analysis of Distribution and
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Levin, S.A., K.D. Kimball, W.H. McDowell, andS.F. Kimball (eds.). 1983.
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APPENDIX A
FRAMEWORK FOR REVIEW OF LEGAL MANDATES
AND ECOLOGICAL ASSESSMENT METHODOLOGIES
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APPENDIX A
TABLE OF CONTENTS
Page
INTRODUCTION A- 1
1.0 Framework for Reviewing Legal Mandates A-2
2.0 Framework for Reviewing Technical Aspects of
Ecological Assessment Methodologies A-2
2.1 Receptor Characterization A-2
2.2 Hazard Assessment A-2
2.3 Exposure Assessment A-6
2.4 Risk Characterization A-6
3.0 Framework for Reviewing Operational
Resource Requirements A-9
LIST OF TABLES
A-l Framework for Reviewing Legal Mandates A-3
A-2 Framework for Reviewing Receptor Characterization A-4
A-3 Framework for Reviewing Hazard Assessment A-5
A-4 Framework for Reviewing Exposure Assessment A-7
A-5 Framework for Reviewing Risk Characterization A-8
A-6 Framework for Reviewing Operational Resource Requirements . A-10
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Review Framework Page A-3
TABLE A-l
FRAMEWORK FOR REVIEWING LEGAL MANDATES
1.0 Legislative Mandate
1.1 Requires ecological risk assessment
1.2 Requires consideration of ecological impacts
1.3 May require consideration of ecological impacts
1.4 Not relevant to ecological impacts
2.0 Executive Order Directive
2.1 Requires ecological risk assessment
2.2 Requires consideration of ecological impacts
2.3 May require consideration of ecological impacts
2.4 Not relevant to ecological impacts
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Review Framework Page A-4
TABLE A-2
FRAMEHORK FOR. REVIEWING RECEPTOR CHARACTERIZATION
1. Type
1.1 Aquatic
1.2 Terrestrial
2. Level
2.1 Individual (e.g., endangered species)
2.2 Population (e.g., game species, keystone species)
2.3 Biotic community (e.g., one .guild, food web)
2.4 Ecosystem
3. Receptor Specificity
3.1 Generic (e.g., generic indicator species)
3.2 Site-specific (e.g., Chesapeake Bay)
4. Temporal Characterization
4.1 Seasonal (e.g., breeding, migration)
4.2 Life cycle
5. Niche (life habit) Characterization
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Review Framework Page A-5
TABLE A-3
FRAMEWORK FOR REVIEWING HAZARD ASSESSMENT
1. Individual/Population Receptors
1.1 Acute criteria (i.e., short duration exposure; e.g., LC5Q)
1.2 Chronic criteria (i.e., longer term exposure; e.g., stunted growth)
1.3 Mechanism of action
1.4 Exposure-response
2. Community/Ecosystem Endpoints
2.1 Criteria
2.2 Exposure-response
3. Toxicity Modifying Factors
3.1 Abiotic conditions/stresses
3.2 Biotic stresses (e.g., competition)
3.3 Behavioral changes
3.4 Bioconcentration :
4. Treatment of Uncertainty
4.1 Generic safety factors
4.2 Ranges
4.3 Statistical extrapolation
5. Hazard Value Derivation
5.1 Apply relative weights to effects
5.2 Choose most sensitive effect
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Review Framework Page A-6
populations, communities) and may consider a variety of receptor-specific
toxicity values (e.g., acute, chronic, dose - response), as well as the general
factors which influence the probability (e.g., toxicity modifying factors) or
interpretation (e.g., hazard value derivation, treatment of uncertainty) of a
toxic response. Ecological assessment methodologies may be characterized and
differentiated to a large degree based on variations in the approaches for
hazard assessment.
Complete evaluation of potential toxicological hazards should include
evaluation of lethal and sublethal endpoints at the individual, population,
community, and ecosystem receptors. Historically, most toxicological data
used in environmental assessments have been measurements of laboratory
population responses, such as LC5Q and MATC values. The majority of current
approaches continue with this emphasis. Community-level toxicity tests (e.g.,
effects on community diversity) have 'not been standardized enough to be widely
used in current methods.
With regard to toxicity values, acute and chronic criteria, mechanisms of
action, and dose-response information have been used or may be used in
ecological hazard assessments and are considered to represent currently -.-.
feasible approaches. Similarly, some assessment methodologies consider
toxicity modifying factors (e.g., abiotic and biotic stresses) and incorporate
estimates of uncertainty.
2.3 EXPOSURE ASSESSMENT
An exposure assessment estimates the chemical-specific concentrations to
which receptors are exposed. As shown in Table A-4, exposure assessments can
encompass multiple exposure pathways, concentration variations, toxicity
modifying factors, and uncertainties surrounding the exposure estimates.
2.4 RISK CHARACTERIZATIOH
The final stage in the ecological assessment process is the integration of
the information derived during receptor characterization, hazard assessment,
and exposure assessment to generate an estimate of impact or risk. Risk
characterization varies with respect to the type of approach/model used to
estimate risk (e.g., qualitative, quantitative); the degree of integration of
multiple stressors or pathways of exposure, and the applicability of the
approach/model to a variety of situations. In addition, models used for risk
characterization vary in the treatment of uncertainty and in the degree to
which they are validated, calibrated, verifiable, and sensitive. Each of
these characteristics is listed in Table A-5.
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Review Framework Page A-7
TABLE A-4
FRAMEWORK FOR REVIEWING EXPOSURE ASSESSMENT
1. Exposure Pathways
1.1 Soil
1.2 Air
1.3 Water
1.4 Sediment
1.5 Food
2. Concentration Estimation Methods
2.1 Spatial variation within/between media (e.g., concentration is a
continuous or step function of distance from source)
2.2 Temporal variation within/between media (e.g., steady state or time
varying contaminant release)
2.3 Toxicokinetics (e.g., bioconcentration of substance)
: '•
3. Characterization of Toxicity Modifying Factors (e.g., D.O., pH,
temperature)
4. Treatment of Uncertainty
4.1 Deterministic model
4.2 Average (i.e., typical case) or maximum (i.e., worst case) estimates
4.3 Probability distribution of concentrations
-------
Review Framework Page A-8
TABLE A-5
FRAMEWORK FOR REVIEWING RISK CHARACTERIZATION
1. Type of Approach
1.1 Qualitative
1.2 Quantitative
2. Degree of Integration
2.1 Contaminants
2.1.1 Single (or indicator) chemical
2.1.2 Multiple chemicals
2.2 Route of Exposure
2.2.1 Single route or dominant route
2.2.2 Multiple routes
2.3 Contributions of Other Stresses
3. Applicability of Model/Approach
3.1 Limited (e.g., specific to a particular situation or locality)
3.2 Broadly applicable/flexible
4. Treatment of Uncertainty
4.1 Reflects uncertainty in measurements
4.2 Reflects biotic and environmental variability
4.3 Estimates probability distribution of effects
5. Can Be Validated
5.1 Has been partially field tested
5.2 Has been completely field tested
5.3 Can be field tested
6. Can Be Calibrated (i.e., adjusted to reflect the environmental situation at
hand)
7. Can be Evaluated for Sensitivity (i.e., changes in output of model
accurately reflect changes of inputs)
8. Is Verifiable
8.1 Model code/behavior accurately represents algorithm
8.2 Algorithms are scientifically reasonable and/or field verified
-------
Review Framework Page A.-9
3.0 FRAMEWORK FOR REVIEWING OPERATIONAL RESOURCE REQUIREMENTS
Our review of operational resource concerns addresses the issue that while
some aspects of methodologies do not affect the scientific approaches of the
model, they may influence the degree to which the model is applied. For
example, a complex computer simulation model may be extremely expensive to
run, and this may limit its use. In addition, complex models may require data
that are not available for a wide range of chemicals or organisms. The
primary operational resource concerns considered are data availability, cost,
and the level of effort and skill required to implement the ecological
assessment. The framework for reviewing operational resource requirements is
given in Table A-6.
-------
Review Framework Page A-LO
TABLE A-6
FRAMEWORK FOR REVIEWING OPERATIONAL RESOURCE REQUIREMENTS
1. Data Availability
1.1 Available (i.e., most receptors, most chemicals)
1.2 Limited availability (i.e., few receptors, few chemicals)
1.3 Can be extrapolated using similar species/populations or chemicals
2. Cost/Level of Effort
2.1 High
2.2 Medium
2.3 Low
3. Level of Skill Required
3.1 High
3.2 Low
-------
APPENDIX B
REVIEW OF LEGAL MANDATES FOR ECOLOGICAL RISK ASSESSMENTS
-------
TABLE OF CONTENTS
Page
INTRODUCTION AND MAJOR CONCLUSIONS B-1
1.0 Statutes and Directives Administered by EPA
1.1 Comprehensive Environmental Response,
Compensation, and Liability Act (CERCLA) B-3
1.2 Clean Air Act B-12
1. 3 Clean Water Act B-12
1.4 Federal Insecticide, Fungicide, and
Rodenticide Act (FIFRA) B-13
1.5 Marine Protection and Research
Sanctuaries Act (MPRSA) B-14
1.6 Resource Conservation and Recovery
Act (RCRA) B-14
1.7 Safe Drinking Water Act (SBWA) B-15
1.8 Toxic Substances Control Act (TSCA) B-15
2.0 Statutes and Directives Administered by Other Federal Agencies,
2.1 Coastal Zone Management Act of 1972 B-17
2.2 Endangered Species Act B-17
2.3 Executive Order 11988, Floodplain Management B-17
2.4 Executive Order 11990, Protection of Wetlands B-17
2.5 Fish and Wildlife Coordination Act -. B-20
2.6 National Environmental Policy Act (NEPA) B-20
2.7 Wild and Scenic Rivers Act B-20
B-i
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Legal Mandates for Ecological Risk Assessment Page B-l
INTRODUCTION AND MAJOR CONCLUSIONS
This appendix presents a review of legal mandates for ecological risk
assessment. The review includes 15 statutes and directives, eight of which
are administered by EPA, with the remaining seven administered by other
agencies within the Federal government. We also examined the legislative
histories for several statutes in order to interpret the intent of specific
provisions within the legislation. It should be noted that a review of
regulations promulgated in the Federal Register was outside the scope of this
project. Hence, specific regulations for ecological risk assessment such as
those promulgated under FIFRA (e.g., 40 CFR 162.11) are not addressed in this
document. The statutes and directives administered by EPA are presented
first, followed by the legislation administered by other Federal agencies.
Protecting the environment or various components of the environment is a
clear common goal or objective of each of the 15 statutes and directives
reviewed. The statutes and mandates*do not describe specifically how
protection of the environment is to be accomplished. Nevertheless, given the
goals and objectives of the statutes and directives, it is arguable that all
may require some consideration of ecological impacts even though none
explicitly requires formal ecological risk assessment. However, all of them
allow, and some appear to actually encourage, conducting an ecological risk'
assessment.
Of the eight statutes administered by EPA, four appear to require
consideration of ecological impacts in support of standards, limitations, and
regulations development. Ecological impact considerations are also required
to support waivers, exemptions, etc., from regulatory standards, limitations,
and requirements, and to support research activities related to the impacts of
regulations or in support of future regulations. These statutes specify the
activities or biota for which a consideration of environmental impact is
required (e.g., propagation of shellfish, fish, etc.). The remainder of the
statutes administered by EPA may require consideration of ecological impact
and would allow the use of administrative discretion to require an assessment
to ensure protection of the environment. In general, these statutes and
directives do not specify activities and biota for which an assessment must be
provided or which methods must be used in performing the assessment.
The Clean Air Act (CAA), Clean Water Act (CWA), the Federal Insecticide,
Fungicide, and Rodenticide Act (FIFRA), the Marine Protection, Research, and
Sanctuaries Act (MPRSA), the Endangered Species Act (ESA), the National
Environmental Policy Act (NEPA), the Wild and Scenic Rivers Act, and Executive
Order 11990 (Protection of Wetlands) require consideration of ecological
impacts. Of these statutes, the CAA, CWA, FIFRA, and MPRSA are administered
by EPA. Each of these statutes contains sections for which consideration of
ecological impacts is required and sections which seem to encourage performing
formal ecological risk assessments. For example, to obtain a modification of
secondary treatment requirements under Section 301(h) of the CWA, a
municipality must first demonstrate that the modification will not result in
interference with maintenance of water quality that ensures the "protection
and propagation of a balanced, indigenous population of shellfish, fish, and
wildlife." Under Section L02(a) of MPRSA, prior to issuing a permit for ocean
-------
Legal Mandates for Ecological Risk Assessment Page B-2
dumping, the Administrator must consider the effects on marine ecosystems,
including the transfer, concentration, and dispersion of waste, the potential
changes in marine ecosystem diversity, productivity, and stability, and marine
species and community population dynamics.
The Comprehensive Environmental Response, Compensation, and Liability Act
(CERCLA), the Resource Conservation and Recovery Act (RCRA), the Toxic
Substances Control Act.(TSCA), the Safe Drinking Water Act (SDWA), the Coastal
Zone Management Act of 1972 (CZMA), the Fish and Wildlife Coordination Act,
and Executive Order 11988 (Floodplain Management) may require consideration of
ecological impacts. Of these, CERCLA, RCRA, SDWA, and TSCA are administered
by EPA. Sections within each of these statutes require that environmental
impacts be considered. Implicitly, this includes ecological impacts. In
addition, each statute provides sufficient administrative discretion and
flexibility to require this consideration. For example, authorization for
the land disposal of hazardous wastes under RCRA Sections 3004(d) and (e) can
be granted only for disposal methods that are shown to be protective of human
health and the environment.
-------
Legal Mandates for Ecological Risk Assessment Page B-3
1.0 STATUTES AND DIRECTIVES ADMINISTERED BY EPA
Table B-l provides a summary of ecological assessment requirements of
eight statutes and directives administered by EPA. A brief description of the
requirements for each statute and directive is provided belov.
1.1 Comprehensive Environmental Response. Compensation, and Liability
Act (CERCLA)
The objective of CERCLA (including amendments made under the Superfund
Amendments and Reauthorization Act of 1986) is to establish a National program
for responding to releases of hazardous substances into the environment.
CERCLA establishes the process for determining appropriate remedial actions at
Superfund sites; remedial actions selected are required to be cost-effective
and adequate to protect public health.
Several sections of CERCLA require that environmental impact assessments
be performed for certain activities, although neither the Act nor its
legislative history specifies methods or techniques for the assessments. The
environment is defined to include natural resources that are under the
exclusive management authority of the U.S. Natural resources are further
defined under CERCLA to include land, fish, wildlife biota, air, water, ground
water, drinking water supplies, and other resources. Therefore, although
there is not an explicit requirement for an ecological assessment, the
definitions provided could be interpreted to require one.
Under the National Contingency Plan, Section 105 of CERCLA, several
provisions require the consideration of the threat of, and subsequent hazards
related to, the release of hazardous substances to the environment. Section
105(a)(8)(A) requires criteria to be established for determining priorities
among releases or threatened releases of hazardous substances for purposes of
taking removal actions. These criteria must be based upon "the relative risk
or danger to... the environment, (including) the potential for destruction of
sensitive ecosystems." Under Section 105(d) of CERCLA, the President can be
petitioned to conduct a preliminary assessment of a release or threatened
release of a hazardous substance. This preliminary assessment must consider
the hazards associated with a release or threatened release posed to public
health and the environment. To add facilities to the National Priorities List
(NPL) pursuant to Section 105(g) of CERCLA, the potential exposure and hazard
of substances to human health and the environment must be considered. Section
121 (added pursuant to the 1986 amendments) of CERCLA defines how to assess
alternative remedial actions and the degree of cleanup required for Superfund
sites. This assessment must account in part for the potential threat to human
health and the environment. Section 301(c) of CERCLA specifies that two types
of procedures be developed to assess damages "for injury to, destruction of.
or loss of natural resources resulting from a release of oil or a hazardous
substance." The first type of procedure must include minimal field
observation. The second type must include protocols for assessing damage on a
case-by-case basis, including identification of the best available procedures
to determine such damages considering factors such as replacement value, use
value, and ability of ecosystem or resource to recover.
-------
Legal Mandates for Ecological Risk Assessment
Page B-4
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TABLE B-l (Continued)
SUMMARY Of THE LEGAL MANDATE POi ECOLOGICAL ASSESSMENT (EPA>
L.gi.l.Hon
Applicable
Section(s)
Activities For Which Ecological Assessment
Mquir«d/««coi««d.
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after date of enactment).
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site and surrounding environment.
Applicant nust carry out a comprehensive monitoring plan
that includes fish and benthic anisial sampling.
Applicant must aiaess analysis of environmental impact on
human health and welfare and marine life.
202(a)<2)
Research on ocean dumping
Comprehensive research on ocean dumping is required to
define and quantify the degradation of the marine envir-
onment and assess the health of the marine environment.
p>
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Legal Mandates for Ecological Risk Assessment
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-------
TABLE B-l (Continued)
SUMMARY OP THE LEGAL MANDATE FOB ECOLOGICAL ASSESSMEKT (EPA)
sr
04
Legislation
Applicable
S«ction(a)
Activities Por Which Ecological AsaeeaaMnt
Required/Recommended
Citation Summary
rt
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(cont inued)
SDUA
TSCA
8002U),(m),
(n),(o),l (p)
lA27(d)
4(a)
Ub)
S(b)
Special watte studies
Development of criteria for the identification of critical
aquifer protection areaa.
Testing to develop data with reapect to health and
environmental effecta of substances or mixtures manu-
factured, distributed in commerce, proceaacd, used, or
diaposed of.
Establishment of standards for the develop*
data for tubatancea or Mixture*
:nt of teat
Notification and submission of teat data required for
manufacture of a new chemical aubatance or processing
any chemical aubatance for a new use.
Exemption from notification require
•arketing purposes.
nts for teat
Each study matt account for the potential danger to huawn
health and the environment from specific waste management
practices.
Environmental costs and benefits have to be considered
during identification of critical aquifer protection areaa.
Required only when insufficient data exists for the sub-
stance or mixture, to determine if an unreasonable risk of
injury to health and environment is present.
Standards for development of test data for substances or
mixtures may be preacribed for:
o Environmental effects
o Chemical characteristics
o Methodologies for determining effects
Test data required must show that the new substance or
new use will not present an unreasonable riak of injury
to health or the environment.
Must show that the msnufacture, processing, distribution
in commerce, use, and diaposal of a substance will not
present any unreasonable risk of injury to health or the
environment.
6(e)
Authorization to manufacture or use PCRs in a manner
other than totally enclosed.
Substance or mixture being manufactured, processed, or
distributed in commerce for export
Must show that manufacture or use will not present any
unreasonable risk or injury to health and the environment.
May require testing (pursuant to .Section A of TSCA)
if unreasonable risk may be present to health and the
environment.
09
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h-1
O
-------
Legal Mandates for Ecological Risk Assessment Page B-ll
Finally, Section 311 of CERCLA requires basic research of hazardous
substances involving evaluation of methods to reduce the amount and toxicity
of hazardous substances.
1.2 Clean Air Act
The purpose of the Clean Air Act is to protect and enhance the quality of
the Nation's air resources through the prevention and control of air
pollution. Although the primary objective of the Clean Air Act is to promote
the public health and welfare, several provisions in the Act require
evaluation of environment impacts. Subchapter I, Part B, Ozone Protection,
encourages studies to assess effects of changes in the ozone layer upon living
organisms and ecosystems. Subchapter I, Part C, Prevention of Significant
Deterioration of Air Quality, requires that the criteria for reclassification
of protected areas and the redesignation of preconstruction requirements for
major emitting facilities include an assessment of environmental effects. The
Act does not specify the methods for or the components of an environmental
assessment.
1.3 Clean Water Act
The objective of the Clean Water Act (as amended by the Water Quality Act
of 1987) is to restore and maintain the chemical, physical, and biological
integrity of the Nation's waters. One of the National goals established by
the Clean Water Act is to attain water quality that provides
for the protection and propagation of fish, shellfish, and wildlife. The
objective and goals of the Clean Water Act are achieved through the control of
discharges of pollutants to surface waters.
The requirements for an ecological assessment under the Clean Water Ace
fall primarily under Title III, Standards and Enforcement, and Title IV,
Permits and Licenses. Under Sections 303 and 304 of the Clean Water Act,
States and EPA are required to develop standards and criteria, considering the
use and value of waters, and must assure the protection and propagation of a
balanced population of shellfish, fish, and wildlife. In addition, States are
required under Section 305 of the Act to provide EPA with a water quality
inventory. This inventory must include an analysis of the extent to which the
quality of State waters provides for the protection and propagation of a
balanced population of shellfish, fish, and wildlife.
Section 307(a) of the Clean Water Act requires EPA to establish a list of
priority pollutants. In adding to or revising this list, EPA must account for
the toxicity, persistence, and degradability of the pollutant, and the
potential presence of affected organisms and the effects on these organisms.
Under Section 311(b), EPA is required to designate as hazardous those
substances that present imminent and substantial danger to fish, shellfish,
wildlife, shorelines, etc. Effluent limitations established for thermal
discharges under Section 316(a) of the Act must assure protection and
propagation of a balanced, indigenous population of shellfish, fish, and
wildlife. Waivers or exemptions from standards established under the Clean
Water Act (e.g., 301(g) and 301(h)) may not be granted unless it is shown that
the variance will not interfere with the maintenance or attainment of water
-------
Legal Mandates for Ecological Risk Assessment Page B-12
quality that assures the protection and propagation of a balanced population
of shellfish, fish, and wildlife.
Under Title IV, EPA is required to establish control over discharges to
the Nation's waters. The criteria for discharges to oceans under Section
403(c) must consider the effects on marine life, including changes in marine
ecosystem diversity, productivity, and stability. Section 404(c) of the Act
requires an assessment of adverse effects on the environment including impacts
to shellfish beds, fishery areas, and wildlife prior to approval of discharges
from dredge and fill activities.
Several provisions under the 1987 Amendments to the Clean Water Act
require an assessment of environmental impacts. Under Section 304(a), EPA
must develop methods, including biological monitoring and assessment methods,
for establishing and measuring water quality criteria for toxic pollutants.
The 1987 Amendments to the Act also establish, under Section 320(b), an
estuary protection program that includes the development of a management
conference to implement the program. The purpose of a management conference
is to collect and assess data on toxics, nutrients, and natural resources
within the estuarine zone to locate the causes of environmental problems.
1.4 Federal Insecticide. Fungicide, and Rodenticide Act (FIFRA) '"
The objective of the Federal Insecticide, Fungicide, and Rodenticide Act
is to prevent "unreasonable adverse effects on the environment" from the
use of pesticides. Although ecological assessments are not explicitly
required, the components for protection of the environment are generally
provided for through definitions provided in FIFRA. Specifically, the
environment is defined under FIFRA to include "water, air, land, and all
plants and man and other animals living therein, and the interrelationships
which exist among these." Unreasonable adverse effects on the environment are
defined under FIFRA as any "unreasonable risk to man or the environment,
taking into account the economic, social, and environmental costs and
benefits."
Section 3 of FIFRA establishes the procedures by which a. pesticide is
registered. Section 3(c) specifies that prior to approval of a. registration,
an assessment of the impacts of the use of a pesticide on the environment must
be performed by the Administrator. Under Section 3(d), the potential for a
pesticide to adversely effect the environment if improperly used is the
primary factor the Administrator must consider when classifying pesticides for
general or specific use. For pesticides not previously registered, the
Administrator can issue an experimental use permit under Section 5 of FIFRA
that would contain conditions for use. The Administrator may require, as a
condition to the permit, studies to detect whether the pesticide under the
permit may cause unreasonable adverse effects on the environment.
In addition to setting forth criteria for pesticide classification and
use, FIFRA specifically calls for assessments of environmental effects of
pesticide use. For example, Section 20(c) of FIFRA requires, as necessary,
monitoring of air, soil, man, plants, and animals to assess incidental
pesticide exposure.
-------
Legal Mandates for Ecological Risk Assessment Page B-13
Finally, Section 25 of FIFRA provides the Administrator with the
authority to prescribe regulations to control the development and use of
pesticides. These regulations must account for the differences in
environmental risk and data for evaluating such risk between agricultural and
nonagricultural pesticides.
1-5 Marine Protection. Research, and Sanctuaries Act (MPRSA1
Title I of the Marine Protection, Research, and Sanctuaries Act
establishes a program to regulate the disposal of waste in the ocean in order
to protect human health, welfare, and amenities, and the marine environment,
ecological systems, and economic potentialities. Section 102 of MPRSA
requires an assessment of the impacts of ocean dumping activities on the
environment prior to issuance of a permit. According to Section 102(a), some
of the factors that the Administrator must consider in issuing permits for
disposal are the effects of the ocean dumping activity on fisheries resources,
plankton, fish, shellfish, wildlife, shore lines and beaches; potential
effects on marine ecosystems; and potential changes in marine ecosystem
diversity, productivity, and stability. Under Section 102(c), the
Administrator can designate recommended dumping sites or times in order to::
protect critical areas. This designation is to be based upon the same
considerations required under Section 102(a), including in part the effects of
the dumping on fisheries resources, plankton, fish, shellfish, etc.
Title II of the Act requires comprehensive research projects on the
effects of ocean dumping on marine environments. Section 202(a)(2) calls for
the assessment of scientific technologies to define and quantify the
degradation of the marine environment, the assessment of the capacity of the
marine environment to receive materials without degradation, and continuing
monitoring programs to assess the health of the marine environment.
Title III establishes the National Marine Sanctuaries Program. According
to Section 303(b), the criteria for determining whether an area will be
designated as a sanctuary include the natural resources and ecological
qualities of an area, such as biological productivity, maintenance of
ecosystems structure, and maintenance of ecologically or commercially
important or threatened species. The present and potential activities that
may also affect the natural resources and ecological qualities of an area must
also be considered.
1.6 Resource Conservation and Recovery Act (RCRA)
The objectives of RCRA are to promote the protection of human health and
the environment and to conserve valuable material and energy sources.
Although neither the Act nor its legislative history define protection of the
environment, there are several provisions under RCRA that could necessitate an
assessment of environmental impacts. Section 1008 requires EPA to develop
guidelines that provide levels of performance of solid waste management
practices that would provide for the protection of public health and the
environment. Requirements to protect the environment are contained in various
subsections of Section 3004, Standards Applicable to Owners and Operators of
-------
Legal Mandates for Ecological Risk Assessment Page B-14
Hazardous Waste Treatment, Storage, and Disposal Facilities. Specifically,
determinations by EPA for the prohibition or regulation of various solid waste
management practices and associated hazardous wastes must account for
protection of the environment. For example, Section 3004(b) prohibits the use
of underground mines for the storage or disposal of hazardous wastes unless
EPA determines that such placement is protective of human health and the
environment.
Under Subtitle H of RCRA, various studies are required to conduct
research, demonstration, and training related to solid waste management
practices in an effort to improve practices and reduce adverse environmental
effects.
1.7 Safe Drinking Water Act (SDWAV
The purpose of the SDWA is to assure that all public drinking water
systems provide safe, high quality water. Although the primary goal of the
SDWA is protection of public health, one provision requires an environmental
assessment. Section 1427 of the Act requires that EPA develop criteria for
the identification of critical aquifer areas that should receive special
protection. In developing the criteria for this identification, the .
environmental costs and benefits must be considered. The specific factors'
associated with the assessment of environmental costs and benefits are not
provided for in the Act itself, nor are they discussed or interpreted in the
legislative history of the Act.
1.8 Toxic Substances Control Act (TSCA)
The objective of TSCA is to protect human health and the environment
through the regulation of the manufacture and use of chemical substances. Th
specific methods that are necessary to assure protection of the environment
are not provided in the Act. Environmental assessments could be required
based upon the definition of environment provided under the Act. The
environment is defined in TSCA to include "water, air, and land and the
interrelationship which exists among and between water, air, and land and all
living things."
One of the primary means of protection established under TSCA is
requiring test data under Sections 4 and 5. This test data must show that the
manufacture, processing, distribution, use, and disposal of substances will
not present any unreasonable risk of injury to health or the environment.
Section 4 of TSCA requires the submission of test data for substances that
have been included on the priority list, established pursuant to Section 4(e).
Section 5 requires notification of the intent to manufacture or process any
chemical substance. For the manufacture of a new chemical substance or new
uses of an existing chemical substance, test data must be submitted along with
the notification. Exemptions for notification to manufacture or process a
chemical substance can be granted under Section 5(h) for test marketing
purposes. However, this exemption can be granted only after it is
demonstrated that the manufacture or use of the substance will not present any
unreasonable risk of injury to health or the environment.
-------
Legal Mandates for Ecological Risk Assessment Page B-15
Section 6(e) of TSCA requires the establishment of regulations to
specifically control the manufacture and use of polychlorinated biphenyls.
Under Section 6(e), the Administrator is authorized to permit the manufacture,
process, distribution, and use of polychlorinated biphenyls in a manner that
is not totally enclosed only if it is shown that the practice will not pose an
unreasonable risk of injury to health or the environment.
-------
Legal Mandates for Ecological Risk Assessment Page B-16
2.0 STATUTES AND DIRECTIVES ADMINISTERED BY OTHER FEDERAL AGENCIES
Table B-2 provides a summary of ecological assessment requirements of
environmental legislation administered by Federal agencies other than EPA. A
brief description of the ecological risk assessment requirements for each
statute and directive is provided below.
2.1 Coastal Zone Management Act of 1972
The purpose of the Coastal Zone Management Act of 1972 is to encourage
and assist States to develop and implement management programs to protect the
land and water resources of the coastal zone, considering in part the
ecological values of the coastal zone. Although ecological assessments are
not specifically required under this Act, it may be appropriate for an
assessment of ecological risks to be performed as part of a coastal management
program established pursuant to the Act.
2.2 Endangered Species Act
The objective of the Endangered Species Act is to conserve the ecosystems
upon which endangered or threatened species depend, and to conserve the
species themselves. In determining whether or not a species is considered-.
endangered or threatened, Section 4(a) requires that the present or potential
destruction, modification, or curtailment of a species habitat or range, and
other natural and manmade factors affecting the existence of the species, be
assessed. Further, under Section 7, any Federal agencies that undertake
activities that may jeopardize the continued existence of an endangered
species must perform a biological assessment to identify the endangered
species which is likely to be affected by the activity. The specific form of
the biological assessment is not specified in the Act.
2.3 Executive Order 11988. Floodplain Management
Executive Order 11988 requires that agencies of the Federal government
take actions to reduce the risk of flood loss; minimize the impacts of flood
loss; minimize the impacts of floods on human safety, health, and welfare; and
restore and preserve the natural and beneficial values served by floodplains.
Although ecological assessments are not specifically required, any major
Federal actions involving floodplains must be accompanied by an environmental
impact statement as required under the National Environmental Policy Act
(discussed further below).
2.4 Executive Order 11990. Protection of Wetlands
Executive Order 11990 requires that Federal agencies take actions to
minimize the destruction, loss, or degradation of wetlands, and to preserve
and enhance the natural and beneficial values of wetlands. According to
Section 5(b), prior to any Federal activity involving wetlands, the
appropriate Federal agency must consider the effect of the activity on the
maintenance of the natural system, including productivity of existing flora
and fauna, species and habitat diversity and stability, hydrologic utility,
and fish, wildlife, timber, food, and fiber resources.
-------
Legal Mandates for Ecological Risk Assessment
Page B-17
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Legal Mandates for Ecological Risk Assessment Page B-19
2.5 Fish and Wildlife Coordination Act
The objective of the Fish and Wildlife Coordination Act is to assist in
developing and protecting all species of wildlife, wildlife resources, and
wildlife habitat. Several sections under this Act appear to require an
ecological assessment. Specifically, Section 2 of the Act requires a report,
for projects involving impounding, diverting, or controlling waters, which
describes the damage to wildlife attributable to the project and measures
proposed for mitigating or compensating for these damages. In addition, the
report has to include an estimation of the wildlife benefits or losses to be
derived from any new project for the control or use of water. Under Section 5
of this Act, the Fish and Wildlife Service and the Bureau of Mines are
authorized to perform investigations to determine the effects of pollutants on
wildlife including the determination .ef standards of water quality for the
maintenance of wildlife.
2.6 National Environmental Policy Act (NEPA)
The objective of NEPA is to preserve and enhance the environmental
quality of the Nation. To ensure that the objective is achieved, all Federal
agencies are required to submit a report, for proposed legislation or major
Federal actions that could have a significant impact on the environment, that
describes the environmental impact, any adverse environmental effects, etc.
The Act does not specify the form of the analysis necessary, nor does it
define environment. However, these studies usually include an analysis of the
impact on soils, geology, hydrology, air quality, and wildlife. The Council
on Environmental Quality, established under Title II, is also required to
conduct investigations, studies, surveys, research, and analyses relating to
ecological systems and environmental quality. The Council must present an
Environmental Quality Report to Congress on an annual basis. This report sets
forth the status of the environment, including the air, marine, estuarine, and
fresh water environments, and forest, dryland and wetland environments.
2.7 Wild and Scenic Rivers Act
The goals of the Wild and Scenic Rivers Act are to preserve selected
rivers in their free-flowing condition, to protect the water quality of such
rivers, and to fulfill other vital National conservation purposes. The Act
designates the initial components (i.e., rivers) of the National Wild and
Scenic Rivers System, and prescribes methods and standards according to which
additional components may be added. Although there is no specific requirement
for an ecological assessment in conjunction with the methods and standards
provided, the stated goals indicate that ecological impacts should be
considered under this Act.
-------
APPENDIX C
ECOLOGICAL ASSESSMENT METHOD SUMMARIES
-------
TABLE OF CONTENTS
Page
Introduction C-l
1. Methods/Assessments Developed Under CERCLA
1.1 CERCLA Type A Natural Resource Damage Assessment Model
for Coastal and Marine Environments (DOI 1987a)...... C-2
1.2 CERCLA Type B Natural Resource Damage Assessment
(DOI 1987b) C-10
2. Methods/Assessments Developed Under the Clean Air Act
2.1 Review of the National Ambient Air Quality
Standards for Ozone: Staff Paper (EPA/OAR 1986) .... C-14
2.2 An Assessment of the Risk of Stratospheric
Modification; Submission to the Science
Advisory Board (EPA/OAR 1987) C-17
3. Methods/Assessments Developed Under the Clean Water Act
3.1 Guidelines for Deriving Numerical National
Water Quality Criteria for the Protection
of Aquatic Organisms and their Uses
(EPA/OWRS 1986) C-22
3.2 An Approach to Assessing Exposure to and
Risk of Environmental Pollutants (EPA/OWRS 1983) C-25
3.3 Permit Writer's Guide to Water Quality-Based
Permitting for Toxic Pollutants / Technical Support
Document for Water Quality-Based Toxics Control
(EPA/OWRS 1987, 1985) C-28
3.4 Biological Criteria for the Protection of Aquatic
Life (Ohio EPA 1987a, 1987b, 1988) C-32
3.5 Niagara River Biota Contamination Project: Fish
Flesh Criteria for Piscivorous Wildlife
(NYS/DEC 1987) C-39
4. Methods/Assessments Developed Under FIFRA
4.1 Standard Evaluation Procedure for Ecological
Risk Assessment (EPA/OPP 1986) C-42
4.2 Computer-based Environmental Exposure and
Risk Assessment Methodology for Hazardous
Materials (Chemical Migration Risk Assessment;
Onishi e_£ al. 1982, 1985) C-47
5. Methods/Assessments Developed Under RCRA
5.1 Potential for Environmental Damage:
Proximity of Mine Sites to Sensitive
Environments (EPA/OSW 1987a) C-52
5.2 Variances from the Secondary Containment
Requirements of Hazardous Waste Tank Systems:
Risk-based Variance (EPA/OSW 1987b) C-54
5.3 The RCRA Risk-Cost Analysis Model (EPA/OSW 1984) C-57
C-i
-------
LIST OF FIGURES
Page
Cl-1 NRDAM/CME Model (DOI 1987a) C- 3
Cl-2 Food Web Model for CERCLA Type A Assessments (DOI 1987a) .. C- 7
C2-1 Generalized Flow Diagram for Assessment of
Stratospheric Modification (EPA/OAR 1987) C-18
C3-1 Schematic Diagram of EPA/OWRS (1983) Approach to
Risk Assessment C-26
C3-2 Biological Indices of Surface Water Quality
in the Scioto River (Ohio EPA 1988) C-35
C3-3 Longitudinal Trends in Biological Indices of Surface
Water Quality in the Cuyahoga River (Ohio EPA 1988) C-36
C4-1 Schematic Diagram of the CMRA Methodology
(Onishe et aj.. 1982, 1985) C-48
it
C5-1 The Ecosystem Damage Function (EPA/OSW 1984) C-58
C6-1 Example of a Fault Tree Analysis (Barnthouse ej£ al. 1986
as used in EPA/OTS 1987) C-64
C7-1 Overlap in EEC and MATC Probability Distributions
(Barnthouse e_£ al. 1986) C-71
C7-2 Percent Response vs. Concentration
(Barnthouse e£ al. 1986) C-74
C7-3 The SWACOM Computer Model (Barnthouse et §1. 1986) C-76
-------
TABLE OF CONTENTS
(continued)
Page
6. Methods/Assessments Developed Under TSCA
6.1 Estimating "Concern Levels" for
Concentrations of Chemical Substances
in the Environment (EPA/OTS 1984) C-61
6.2 Ecological Risk Assessment in the Office of
Toxic Substances, Problems and Progress
(EPA/OTS 1987) C-63
7. Other Methods
7.1 Users Manual for Ecological Risk Assessment
(Barnthouse e_£ al. 1982, 1986) C-67
7.2 Regional Ecological Assessments: Concepts,
Procedures and Application (Ballou §_£ a\. 1981) C-79
7.3 Computer Simulation Models of Assessment of Toxic
Substances (Eschenroeder e_t al,. 1980) C-87
7.4 Unfinished Business: A Comparative Assessment
of Environmental Problems. Appendix III,
Ecological Risk Work Group (EPA/OPPE 1987) C-89
C-ii
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LIST OF TABLES
Page
C4-1 Wildlife and Aquatic Organism Data Requirements for
the Standard Evaluation Procedure for Ecological Risk
(EPA/OPP 1986) C-44
C4-2 Selected Bibliography for the Chemical Migration
Risk Assessment methodology C-49
-------
Method Summaries Page C-l
INTRODUCTION
In this appendix, we present a review of twenty ecological assessment
methodologies that have been developed for use by one or more Federal or State
agencies. This is not a comprehensive list of available ecological assessment
methodologies. The methodologies included in our review are representative of
those used for three general types of applications: setting environmental
criteria, setting priorities, and risk characterization for purposes of risk
management.
We describe how the methodologies characterize ecological targets,
hazard, exposure, and risk. The reviews are organized by the legal mandates
under which the methodologies were developed. Methods developed under CERCLA
are presented first, followed by those developed under the Clean Air Act,
Clean Water Act, FIFRA, RCRA and TSCA. Several other methodologies, not
developed under any specific legal mandate, also are included in this review.
-------
CERCLA Type A Damage Assessment DOI 1987a Page C-2
1.0 METHODS/ASSESSMENTS DEVELOPED UNDER CERCLA
1.1 CERCLA TYPE A DAMAGE ASSESSMENT FOR COASTAL AND MARINE ENVIRONMENTS (DOI
1987a)
1.1.1 Introduction
The Department of the Interior (DOI) has developed procedures for
assessing damages to natural resources from a discharge of oil or a release of
a hazardous substance compensable under either CERCLA or the Clean Water Act.
Under CERCLA, states or the Federal government can assert claims for
compensation for natural resource damage or injury following discharge of oil
or hazardous substance release. The Natural Resource Damage Assessment Model
for Coastal and Marine Environments (NRDAM/CME) was developed for damage
assessment following a single, relatively small (type A) release in coastal
and marine environments. The model is an interactive computer program that
generates a measure of damages based on information regarding the nature,
date, and location of a hazardous material release. Minimal field
observations are required.
In the NRDAM/CME model, damages are defined as diminution in the use •-
value of natural resources resulting from injuries caused by a discharge or
release. Damages are equal to the difference in the i,n situ resource values
pre- and post-incident.
As Figure Cl-1 illustrates, the model comprises three major submodels and
data bases. The physical fate and transport submodel simulates the spread of
contaminants from a discharge source over time and space. The biological
submodel estimates both direct and indirect effects on plants and animals.
(Direct effects include contaminant-induced mortality; Indirect effects
include lost productivity in higher trophic levels as a consequence of
mortality in lower trophic levels and reduced recruitment of larvae and
juveniles.) The economic submodel estimates damages to commercial and
recreational fisheries. Damages to public beaches, waterfowl, shorebirds,
seabirds, and marine mammals are also estimated. Biotic effects such as
reduced growth rates, lost reproductive potential, increased susceptibility to
other biotic stresses (e.g. disease, parasitism), or competitive relationships
between species are not addressed.
1.1.2 Description of Method
Receptor Characterization
Receptor characterization under the NRDAM/CME model is limited to coastal
and marine environments. Within these environments, a variety of marine and
coastal bioraes are represented. Coastal waters are divided into ten regions
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CERCLA Type A Damage Assessment DOT 1987a
Page C-3
FIGURE C14
NRDAM/CMB MODEL
(Source: DOI 1987a)
USM
snu. rvpe. town ON. OATI.
HAllTAf CUSSiriCAriQN, MACM/
HUNTINC/riSWHC
-------
CERCLA, Type A Damage Assessment DOI 1987a Page C-4
or provinces which are further divided according to tidal position (i.e.,
inter tidal/sub tidal) and bottom type (e.g., rock, sand, coral). For each of
the resulting 91 ecosystem types, there are four categories of data: (1) adult
biomass by species; (2) larval numbers by species category, (3) mortality and
growth parameters by species category; and (4) primary and secondary
productivity values.
Biological receptors are divided into categories: six fish species
groups, three invertebrate groups, three bird groups, and one mammal group.
The fish groups include anadromous, planktivorous , piscivorous, demersal, and
semi-demersal fish and top carnivores. Three life stages are considered for
fish and invertebrates: egg/larval, juveniler- and adult. Shorebirds,
seabirds, waterfowl, and fur seals are represented by two life stages (adults
and young). Primary producers (e.g., phytoplankton, benthic micro- and
macroalgae, and higher plants) and zoop lank ton are also considered.
Seasonally dependent parameters considered in receptor characterization
include temperature, adult and juvenile biomass by species group, abundance of
birds and mammals (and their young), number of larvae, and primary
productivity.
Hazard Assessment
The model includes toxicological data for 469 substances. Hazard
assessments for fish, invertebrates,' zooplankton, and phytoplankton are
conducted using LC5Q and £€50 data. A hazard value for each species group is
derived by dividing the acute toxicity value by a safety factor of 100 to
estimate a no-effect level. This approach assumes that the same dose-response
relationship holds for all hazardous substances. Hazard values were derived
in this manner for five organism groups: (1) adult and juvenile fish, (2)
adult and juvenile benthic invertebrates, (3) larvae of fish and benthic
invertebrates, (4) zooplankton, and (5) phytoplankton and other primary
producers. For fish and invertebrates, the acute LC50 values in the data base
were standardized to 25 degrees Centigrade and 96 hours of exposure. Sources
of uncertainty surrounding the hazard values are not addressed.
This hazard assessment methodology includes consideration of the effects
of temperature and duration of exposure (if less than 4 days) on the
population response to a particular toxicant concentration. For exposures
lasting more than 4 days, the 4-day LC^Q is used.
Toxicant -induced effects on primary productivity also are considered in
the hazard assessment. Decreases in primary productivity are estimated from
£€50 data using a log-normal toxicity model. Decreases in productivity are
assumed to be independent of exposure duration.
Potential hazards to other organisms are based on more limited data.
Birds and fur seals are assumed to be affected only by those substances that
form surface slicks (petroleum products in most cases). Based on available
data, it is assumed that a fixed proportion of birds and seals that are oiled
(according to the exposure submodel) will die (38% and 63%, respectively).
-------
CERCLA Type A Damage Assessment DOI 1987a Page C-5
The uncertainty around these estimates, either as representative means or
environmentally dependent variables, is not addressed.
Exposure Assessment
The physical fates submodel estimates the distribution of a contaminant
on the sea surface, in the water column (partitioned as upper and lower), and
in the sediments. The user must enter specific information concerning the
spill (e.g., type and amount of substance spilled, regional province, and
bottom type). The chemical data base of 469 compounds contains empirical or
estimated values for parameters used in the fate and transport model (e.g.,
solubility, degradation rate in water and sediments, and viscosity).
The model is designed to simulate the fate of pure hazardous substances,
crude oils, and petroleum products. 'Crude oil and petroleum products are
represented by four components: (1) low molecular weight aromatics (50/50
mixture of benzene and toluene); (2) medium molecular weight aromatics (100-
160 g/mole); (3) an insoluble but volatile portion; and (4) a residual
fraction which is neither soluble nor volatile.
:t
The time course of events is modeled using incremental time steps scaled
with respect to the size of the spill. Spatial variation is similarly modeled
with a variable-sized horizontal grid. The space in which the toxicant
concentration is estimated to be greater than the toxicity threshold is
divided into ten spatial volumes horizontally in each of three layers (upper
water column, and lower water column, if appropriate; and sediments).
The concentration estimation methods consider several processes affecting
contaminant movement within and between compartments. Contaminants on the
ocean surface spread under the influence of wind, currents, temperature,
viscosity, evaporation, thickness of surface slick, and dissolution into the
water column. Contaminants within the water column are transported
horizontally by current movement and dispersion. Contaminants with a specific
gravity greater than seawater are modeled by a convective descent algorithm.
Contaminants that reach the seafloor enter the sediments via bioturbation
(e.g., bottom fish foraging). The particulate and interstitial concentrations
of the contaminant are modeled separately. Dissolution of contaminants from
the sediment into the water column is modeled. Degradation is modeled in both
the water column and sediments. For substances spilled in the intertidal
area, the model calculates the area impacted and a removal constant which
accounts for dissolution, evaporation, and degradation. All processes are
deterministic except for vertical dispersion which is simulated as a random
walk. Other sources of uncertainty are not developed or propagated.
Risk Characterization
The biological effects submodel of the NRDAM/CME model combines the
quantitative exposure assessment model results with the biological data base
and indicates the numbers and biomass of organisms affected. The submodel
estimates both short- and long-term losses to each of the biota (receptor)
groups. Short-term losses include mortality that is a direct result of
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CERCLA Type A Damage Assessment 001 1987a Page C-6
exposure to toxicant concentrations above threshold. Long-term losses include
lost recruitment of larvae and juveniles into the breeding populations as well
as reduction in productivity at higher trophic levels as a result of lost
primary and secondary production. Other, more complicated biotic
relationships, such as competitive release and density-dependent predator-prey
population effects, are not included in this model because of the lack of a
strong empirical and theoretical framework for such analyses.
Single contaminants or petroleum products that are compartmentalized into
four subgroups can be modeled. Exposure pathways include the water surface,
water column, and sediments but not the transfer of toxicants up the food
chain. The environmental injury submodel is sufficiently simple to be
generally applicable across coastal and marine environments. It is
sufficiently complex to incorporate several levels of effects.
Short-term losses are calculated in the acute toxicity submodel. This
model estimates the fraction of each population (organism class) exposed to a
given concentration that die based on the appropriate LC5Q value, temperature,
and duration of exposure (if less than 4 days). The biomass of each category
of organism is obtained from the biological data base. As time progresses,
more organisms are exposed to the expanding contaminant plume. In addition',
fish are assumed to swim into and out of the plume, thus exposing even more
individuals. Behavioral avoidance of the contaminant plumes is not
considered.
Long-term losses depend upon (1) lost primary and secondary production
and (2) lost recruitment of larvae and juveniles. Reduced primary production
is estimated for phytoplankton and macroalgae, when appropriate, using EC^Q
information. The reduction is passed on to two classes of secondary
production: zooplankton and benthic invertebrates. Direct losses of secondary
producers from toxic effects are also calculated, and the larger of the two
types of losses (indirect through reduced primary production and direct
mortality) is used. This assumes that the growth of zooplankton will either
be food-limited or toxicant-limited, but not both.
Lost primary and secondary productivity is translated into lost yield for
each animal species category via reduced food supply using a simple food web
model which is illustrated in Figure Cl-2. The model is based on simplifying
assumptions which are applicable to all environments.
1 Food chain efficiencies are assumed to be 20 percent. The food web
model assumes that the share of each prey item that a predator consumes is
proportional to its biomass relative to all other predators of that prey.
Larvae of fish and benthic invertebrates are assumed to feed entirely on
zooplankton and to be food-limited.
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CERCLA Type A Damage Assessment DOI 1987a
Page C-7
FIGURE Cl-2
FOOD WEB MODEL FOR CERCLA TYPE A ASSESSMENTS
(Source: DOI 1987a)
Pp. PRIMARY
PRODUCTION
ANAOROMOUS
FISH
(I)
PISCIVOROUS
FISH
(3)
TOP
CARNIVORES
(4)
BENTHOS
P..to. P.
(includes 7* 8)
DEMERSAL
FISH
(S)
SEMI-
DEMERSAL
FISH
(6)
PLANKTIVOflOUS
FISH
(2)
WATERFOWL
AND
SHOREBIRbs
(II, 13)
-------
CERCLA Type A Damage Assessment DOI 1987a Page C-8
Productivity and rates of transfer of energy from lower to upper trophic
levels are province-specific and vary by estuarine or marine environment, by
bottom type, and by season.
Reduction of larval populations, either directly or indirectly, is
translated into long-term catch losses using a fisheries model. Density-
dependent population dynamics are not incorporated. Other adverse effects
modeled include loss of waterfowl for hunting, losses of birds for
recreational purposes (i.e., bird watching), lost fur seal production, and
closed beaches.
Once measures of specific categories of injury have been estimated, the
economic damages submodel places a dollar value on each type of loss. Damages
to lower trophic biota, for which use or market values do not exist, are
translated through the food web into reductions in species of monetary value.
Region-specific valuations of commercial fish species are included. The lost
iB situ value of each biological compartment is calculated separately for each
year.
Although sections of the model might be amenable to field validation, it
is unlikely that the effort involved would be appropriate. A more cost- :
effective approach would be through verification of the algorithms and
sensitivity analyses. The sensitivity analyses presented by the authors
illustrated the following. Economic damages increased linearly with the
amount of contaminant spilled. Damages from the same size spill depended upon
season and province, with winter Arctic coastal spills three orders of
magnitude less costly than spring spills in the California estuarine
environment. In addition, the economic damages varied dramatically with the
physical/chemical properties of the substance spilled.
1.1.3 Operational Resource Requirements
The NRDAM/CME model can be run on a personal computer, and can be
obtained on a series of four diskettes along with the documentation from NTIS.
These diskettes include all associated data bases for the ecological regions
and environments and complete data bases for 469 chemical compounds. With a
minimum of information concerning an actual spill, the model is very easy to
run. The high cost and effort required to develop a model of this type has
already been invested. It is not known if the NRDAM/CME model will be updated
in future years. The computer time required to analyze a single'spill depends
in large part on the size of the spill and can range from 10 minutes for small
spills (e.g., < 10 metric tons) to over an hour for large spills (e.g'( > 500
metric tons).
1.1.4 Summary
The NRDAM/CME modeling approach is based entirely on acute toxicity data
for five general classes of organisms for each compound in the chemical data
base. From these data and a simple food-web model, population effects for
several categories of animals and plants are determined. When the empirical
data indicate that an effect is highly variable in the real world, a "best"
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CERCLA Type A Damage Assessment DOI 1987a Page C-9
case estimate rather than a worst case estimate is selected as the input
value. All output parameters are also represented by single values. No
attempt is made to incorporate the uncertainty in either value measurements or
in the real world. Considering the sizable data bases and the number of
estimations, extrapolations, and assumptions, providing an analytic
characterization of uncertainty from all sources would be difficult.
The NRDAM/CME methodology is a relatively sophisticated approach to
ecological damage assessments for coastal and marine natural resources. In
the context of providing a framework for Type A damage assessments for CERCLA,
the final endpoints of concern are translated into economic losses.
Nonetheless, a wide diversity of endpoints are considered, analyzed, and
presented along with the final monetary evaluations.
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CERCLA Type B Damage Assessment DOI 1987b Page C-10
1.2 TYPE B HAIOBAL RESODRCK DAMAGE ASSESSMENT (DOI 1987b)
1.2.1 Introduction
The Department of the Interior (DOI) has developed procedures for
measuring damages to natural resources resulting from a discharge of oil or a
release of a hazardous substance compensable under either CERCLA or the Clean
Water Act. The Type B damage assessment procedure is intended to provide
guidance for conducting site-specific assessments based on data collected in
the field. The procedure is different from the other assessment methodologies
included in our review because it focuses on field measurements of impacts
that have already occurred, rather than the probability that impacts will
occur. Therefore, hazard and exposure assessments, typical components of
ecological risk assessments, are not conducted, nor is there a risk
model/method to combine hazard and exposure information. Nevertheless, the
damage assessment procedures provide some conceptual approaches that may be
potentially applicable to ecological risk assessment.
Type B natural resource damage assessments involve three major steps:
establishing that an injury has occurred and that the injury resulted from a
discharge or release; quantifying the effects of the discharge or release; and
determining the damage. Injury is defined as a measurable adverse change in
the chemical or physical quality or viability of a natural resource. Damage
is defined as the amount of money sought by the Federal or state agency acting
as trustee for the natural resource to compensate for the loss. Natural
resources include surface water, ground water, geological resources (soil,
rocks, and minerals) and biological resources. Injury is defined uniquely for
each natural resource and is documented following collection of field or
laboratory data and statistical analysis.
1.2.2 Description of Method
Receptor Characterization
Potential receptors include representative components of the entire
biosphere; both the biotic (aquatic and terrestrial) and abiotic components
(air, land, water) of ecosystems are considered. Assessments of biological
injury are conducted at the population level.
Hazard Assessment
As discussed previously, the damage assessment does not have a "true"
hazard assessment procedure. However, the definition of injury for each
natural resource may be useful in characterizing hazard assessment endpoints
for ecological risk assessments.
Injury to surface water and ground water occurs when the water
concentrations of substances are in excess of Federal or State drinking water
standards or water quality criteria, or are sufficiently elevated to cause
damage to other natural resources. Also, an injury has occurred if the
substance's concentration in bed, bank, or shoreline sediments is sufficient
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CERCLA Type B Damage Assessment DOI 1987b Page C-ll
to cause the sediment to exhibit characteristics of hazardous waste (as
defined under RCRA).
Injury to air occurs if the concentrations of emissions are in excess of
standards for hazardous air pollutants established by Federal or State
regulations or if the concentrations are sufficient to cause injury to other
natural resources.
Injury to geologic resources occurs if the concentration of a substance
in the geologic resource is sufficient to cause damage to other natural
resources or is sufficient for the materials in the geologic resource to
exhibit the characteristics of hazardous wastes (as defined under RCRA). Also
an injury has occurred if the concentration of a substance is sufficient to
alter soil pH, salinity, or water-holding capacity beyond prescribed levels,
to impede soil microbial respiration and soil carbon mineralization, or to
cause a toxic response in plant or soil invertebrates.
A biological resource is injured if it or its offspring have undergone at
least one of the following changes in viability: death, disease, behavioral
abnormalities, cancer, genetic mutations, physiological malfunctions, or
physical deformations. '• •
The method for determining injury.is limited by the capability of the
method to demonstrate a measurable biological response. According to the
criteria established by DOI, only certain procedures within the six injury
categories are sufficient to demonstrate an injury to a biological resource.
These are discussed briefly below by injury type. In all cases, measured
responses in injured populations must be statistically different from control
populations.
Death. Death can be measured in both fish and wildlife. Three
acceptable measures of death in fish are fish kills, in situ bioassays using
cultured or indigenous fish, and laboratory tests of acute or subchronic
toxicity. The only acceptable approach to document oil or hazardous substance
related death in wildlife is a brain cholinesterase inhibition of at least 50%
in field populations compared with control populations. Lesser inhibition,
but in excess of 20%, can be considered contributory to death provided that
such inhibition is statistically separable from controls.
Disease. Acceptable approaches for documenting disease injury in fish
are limited to fin erosion and lesions on fish tissue. Acceptable approaches
for documenting disease in wildlife populations are not available.
Behavioral abnormalities. Procedures adequate to document behavioral
changes in fish are limited to laboratory studies of behavioral avoidance in
species representative of site conditions and observations of clinical
behavioral signs of toxicity. For wildlife, injury has occurred if there is
statistically significant increase in observations of clinical behavioral
abnormalities in field populations versus controls, or if two or more
specimens display behavioral abnormalities while none do in the control
populations.
-------
CERCLA Type B Damage Assessment DOI 1987b Page C-12
Cancer. Statistically significant increases of liver or skin neoplasia
are sufficient to document cancer injury in fish. Injury determination in
fish can be documented via field sampling of natural populations or bioassays
using water from the potentially affected area and standard laboratory assay
species. No acceptable procedures have been identified to document cancer
injury in wildlife populations.
Physiological malfunctions. Altered physiologic parameters accepted as
adequate documentation of physiologic malfunctions are reduced reproduction in
fish; and egg shell thinning, reduced reproduction, reduction in brain ChE,
and decreased delta amino levulinic acid dehydratase (delta ALAD) in birds and
mammals.
Physical deformations. Injury from physical deformations can be
documented in wildlife if there is an increased incidence of external
malformations, skeletal malformations', internal whole organ and soft tissue
malformations, or histopathological effects between suspect polluted locations
and control areas. No such measures were identified as adequate for
documenting physical deformation ini in fish.
Exposure Assessment
An exposure assessment is not included in the damage assessment.
Although the regulations stipulate that exposure pathways must be delineated,
specific procedures to document a pathway are not provided.
Risk Characterization
This method does not include a risk model/method. However, procedures to
quantify damages by establishing a dollar value for the injury have been
outlined. Quantification of injury to biological resources can be based on
changes in populations (often quantified by numerical counts of injured
individuals) or on changes in ecosystems or habitats. The Fish and Wildlife
Service has proposed the use of habitat evaluation procedures (HEP) to
quantify changes in fish and wildlife habitats. HEP is an accounting
procedure for tracking an integrated measure of habitat quality and surface
area of habitat. The integrated measure is called a Habitat Unit (HU). The
number of HUs is defined as the product of the Habitat Suitability Index (HSI)
and the area of available habitat. HSIs and HUs are calculated for certain
"evaluation species". The procedure can be used to quantify reductions in
habitat quality between a polluted area and a control area.
1.2.3 Operational Resource Requirements
The damage assessment procedure is a relatively labor- and cost-intensive
approach because field sampling and/or laboratory testing and economic
analyses are required. Theoretically, selected aspects of this approach could
be modified and applied to ecological risk assessments. However, the
operational resource requirements of a modified approach cannot be evaluated
at this time.
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CERCLA Type B Damage Assessment DOI 198 7b Page C-13
1.2.4
The 001 Type B damage assessment regulations provide procedures for
injury determination. These procedures, however, are not ecological risk
assessment procedures, and are therefore limited in their applicability for
risk assessment. However, compilation of data from Type B damage assessments
could provide valuable case-study data for risk assessment. In addition, the
method does offer some unique approaches to define ecological injury. For
example, under these procedures receptors include both biotic and abiotic
ecosystem components. Many ecological risk assessment methodologies consider
only how changes in the abiotic component can impact the biotic components of
the ecosystem. The approach offered by DOI is unique in that abiotic
ecosystem components are recognized as important resources irrespective of the
associated biotic components.
The DOI damage assessment approach also considers the interactions
between the abiotic and biotic components. Consideration of interactions
between system components can loosely be considered an ecosystem approach.
Although other approaches do consider ecosystem effects, implementation of the
approach is not common. •'•
-------
Ozone Staff Paper EPA/OAR 1986 Page C-14
2.0 METHODS/ASSESSMENTS DEVELOPED UNDER THE CLEAN AIR ACT
2.1 REVIE0 OF THE NATIONAL AMBIENT AIR QUALITY STANDARDS FOR OZONE:
STAFF PAPER (EPA/OAR 1986)
2.1.1 Introduction
The Office of Air Quality Planning and Standards (OAQPS) reviewed the
secondary National Ambient Air Quality Standard (NAAQS) for ozone in their
staff paper dated March 1986. The secondary standard is intended to be
protective of both symptomatic health effects in humans and human welfare,
including impacts on vegetation and natural ecosystems. Of particular concern
are the adverse effects of ozone on plants. OAQPS identified short-term and
long-term ambient ozone concentrations associated with low-effect levels for
plants, and then evaluated whether certain criteria (e.g., the primary NAAQS
for human health effects) would adequately protect agricultural plants and
forests.
The approach used by OAQPS was not developed as a generic risk assessment
methodology. We discuss it here because it is an example of the use of
ecological risk assessment, and can in fact be used for other chemicals. '
2.1.2 Description of Method
Receptor Characterization
The assessment focused on terrestrial plants, particularly crops and
forest species, although aquatic systems were discussed briefly. Although
OAQPS summarized the effects of ozone at the individual, population, and
community levels, the exposure and risk assessments emphasized effects
occurring at the population level.
Hazard Assessment
In the hazard assessment, concentrations of ozone that are toxic to
plants were identified. Both qualitative and quantitative assessments were
conducted. The qualitative assessment included descriptions of toxic
endpoints and toxic concentrations at the ecosystem level (e.g., effects on
aquatic systems bordering damaged forests), as well as at the population and
individual levels (e.g., decreased photosynthesis). Response modifying
factors such as water stress, nutrient deficiency, and pest and disease
interactions also were discussed.
In the quantitative assessment, the hazard evaluation was limited to
population-level effects on agricultural crops because of data limitations.
Hazard was expressed in terms of air concentrations associated with phytotoxic
responses. Hazard was evaluated for two concentration averaging periods: (1)
a short-term peak expressed as a number of occurrences above a given level
(multiple exceedances) and (2) a long-term average (expressed over a growing
season of three months). For the short terra assessment, a peak criterion of
0.08 ppra was selected because concentrations of this level or higher are
-------
Ozone Staff Paper EPA/OAR 1986 Page C-15
definitely injurious to sensitive plant species when they are exposed for even
short periods of time, usually less than one hour (Garner 1988). White pine
has been used as an indicator species, injury being determined by visible
foliar damage and growth measured by tree ring number and width. The long-
term average criterion was derived primarily from yield-reduction experiments
on crops. The protectiveness of the long-term standard to forest species was
addressed qualitatively. Damage as used in the assessment is a chronic
endpoint because it occurs over a lifetime (3-month growing season for crops,
or tens of years for trees).
In the long-term hazard assessment, a 10% reduction in yield was used as
the effect level of concern. The 10% level was chosen because it can be
detected with some degree of confidence and is generally considered a
significant effect. Long-term ozone concentrations associated with a 10%
reduction in yield (0.04 to 0.05 ppm. ozone) were determined from dose-response
curves developed from experiments conducted as part of the National Crop Loss
Assessment Network (NCLAN). This range of concentrations was then compared
with data from greenhouse and controlled environment studies and with ambient
air-exposure studies using crop species. The final range of concentrations
associated with a 10% reduction in yield (0.04 to 0.06 ppm ozone) was
presented as a 7-hour 3-month average. The 7-hour component of this average
was chosen because it may correspond to the period of greatest plant
sensitivity to ozone (0900-1600 hr). The three-month component might not be
applicable to deciduous forests, and is not applicable to coniferous forests.
Moreover, repeated, short-term, high-level concentrations can be more injurious
than continuous low-level ozone concentrations that represent the same
"average" concentration (Garner 1988).
Exposure Assessment
Data from EPA's SAROAD (Storage and Retrieval of Aerometric Data) data-
base system were used to summarize current ozone concentrations in the United
States and to predict future concentrations. Current concentrations of ozone
were compared to the peak concentration of 0.08 ppm proposed in the hazard
assessment as a short-term criterion based on phytotoxicity. Data gathered
from 1982 to 1984 indicated that average hourly ozone concentrations in
agricultural areas exceeded 0.08 ppm on the average 119 times each year. In
remote areas, average hourly ozone concentrations exceeded 0.08 ppm on average
37 times each year. A series of longer-terra averages were also presented.
The SAROAD ozone concentration data were also used to predict ozone
concentrations that would result if the human-health-based peak criterion of
0.12 ppm were attained. Assuming peak ozone concentrations of 0.12 ppm, the
percentage of days with hourly maxima exceeding 0.08 ppm were calculated, and
the 7-hour, 3-month average concentration was estimated on the basis of
significant (but weak) correlations between the second-highest daily maximum
concentration, long-term concentrations and short-term peaks. Assuming a
second-highest daily maximum of 0.12 ppm (corresponding to the human health
criteria), there is a 10% chance that 32% or more of the days in a 3-month
growing seasons will exceed a 7-hour, 3-month average of 0.08 ppm. A 7-hour,
3-month average concentration of 0.052 ppm was predicted with an upper bound
(90%) estimate of 0.063 ppm.
-------
Ozone Staff Paper EPA/OAR 1986 Page C-16
Risk Characterization
Risk to terrestrial plants was evaluated qualitatively and focused on the
risk associated with the human health-based peak criterion of 0.12 ppm. Both
the daylight-hours seasonal mean and 90th percentile concentrations calculated
based on attainment of the 0.12 ppm peak level were compared with the range of
concentrations associated with effects. The degree of short-term risk to
terrestrial plants was addressed by evaluating the percentage of days expected
to exceed 0.08 ppm if the 0.12 ppm criterion were attained. Finally, the
large uncertainty in the correlation between short-term and long-term ozone
concentrations (coefficients of determination ranged from 0.3 to 0.4) was used
to support the proposal of a long-term ozone standard in addition to the
short-term standard.
The assessment has been partially calibrated since the hazard assessment
was based in part on effects using ambient ozone levels and on field studies.
Quantitative validation could be conducted using charcoal filtered ambient air
studies.
2.1.3 Operational Resource Requirements
Data on the toxic effects of ozone were available for many major crop
species and a few natural communities. The experimentation required to
develop exposure-response curves for plant species was extensive and
expensive, as was the data collection done to develop the statistics on
ambient ozone concentrations. Therefore, application of this type of approach
for ecological assessments of other chemicals is likely to be costly and labor
intensive. Once these data have been developed, however, the cost and level
of effort needed to estimate risk is low, and the assessment is
straightforward.
2.1.4 Summary
The use of an extensive monitoring effort and exposure-response
characterization is a powerful and useful combination. Limitations of this
approach include the extrapolation from controlled experiments to plants
growing in the field, and the emphasis on yield reduction as the toxic
endpoint of concern. Strengths of this procedure are that it uses exposure-
response modeling and considers temporal characteristics of the receptor.
Also, community and ecosystem level effects as well as response modifying
factors were qualitatively assessed.
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Stratospheric Modification EPA/OAR 1987 Page C-17
2.2 AH ASSESSMENT OF THE RISK OF STRATOSPHERIC MODIFICATION (EPA/OAR 1987)
2.2.1 Introduction
The purpose of this probabilistic risk assessment was to aid the EPA
Administrator in determining the need for additional controls on the releases
of substances that may affect the stratosphere. The report assessed the
likelihood that:
(1) different human activities would result in climate modification or
in an alteration of the stratosphere's ability to filter ultraviolet
radiation, and
(2) changes in ultraviolet radiation (UV-B) or climate due to
modifications in column ozone or stratospheric water vapor would
have detrimental effects on -public health or welfare.
This review addresses only the ecological components of the overall
assessment. Two physical effects of stratospheric modification were included
in the ecological assessment: increased UV-B, and climate change/sea level
rise. The general approach for assessing ecological risks from the two
physical effects was similar. A deterministic model was used to estimate ••'
emissions and ozone depletion. Additional models were used to estimate the
increase in UV-B and temperature attributable to the emissions. Exposure-
response data or qualitative information was then used to estimate ecological
effects. A generalized flow diagram is shown in Figure C2-1.
2.2.2 Description of Method
Receptor Characterization
Increased UV-Q. Both aquatic and terrestrial receptors were considered
under this approach. Although community -level endpoints were considered
qualitatively, quantitative assessments were limited to the population level,
primarily because of the paucity of exposure -response data for higher level
effects. Single species were chosen as receptors for the quantitative
assessments: soybeans in terrestrial systems and anchovies in aquatic
systems. No attempt was made to generalize from these species to other
species or levels of organization. Temporal characteristics of northern
anchovies were discussed and incorporated into estimates of risk.
Change/Sea Level Rise. The climate change assessment primarily
addressed natural plant communities. Coastal wetlands, particularly those in
Louisiana, Florida, Delaware, New Jersey, and Maryland, were identified as
"receptors" of concern for potential effects of sea level rise.
-------
Stratospheric Modification
EPA/OAR 1987
FIGURE C2-1
Page C-18
GENERALIZED FLOW DIAGRAM FOR ASSESSMENT OF
STRATOSPHERIC MODIFICATION
(Source: EPA/OAR 1987)
Year
Loop
Input
Files
Production
Scenarios
Module
T
Policy
Alternatives
Module
Emissions
Module
r
I
I
I »
Atmospheric
Science
Module
I
Effects
Module
I
Summary
File
^_^^
Report
Writer
Program
Output
Re
^_^x^
-------
Stratospheric Modification EPA/OAR 1987 Page C-19
Hazard Assessment
Increased UV-B. For the qualitative assessment of hazard to aquatic
organisms, effects to a wide variety of species, including fish larvae and
juveniles, shrimp larvae, crab larvae, copepods, and aquatic plants, were
discussed; however, effects on planktonic organisms were emphasized. Most of
the quantitative data available addressed anchovy mortality, but sublethal
effects such as reduction in fecundity were also discussed. Temporal
attributes of the organisms, particularly diel vertical movement of zoo- and
icthyoplankton, were recognized as important uncertainties. Tolerance and
avoidance mechanisms were also discussed. The most complete exposure-response
data were available from laboratory tank experiments with anchovy larvae.
These data were reinterpreted to include scenarios for vertical movement
during the day by the larvae.
The assessment of effects on terrestrial organisms was limited to effects
on plants. Little information was available for natural vegetation. Toxicity
modifying factors such as mineral deficiency and water stress were discussed.
Although effects on yield were emphasized, effects on yield quality and
community-level effects, such as changes in plant competition and increases in
disease and pests, were also discussed. The most complete exposure-response
data were available for soybeans. The extrapolation from controlled
experiments to effects in the field was recognized as a particularly
significant source of uncertainty, but was not factored into the exposure-
response quantification.
C14.pa.pfii, Change/Sea Leve], Rise. The effects of climate change on plants
were addressed using several methods. Paleovegetational records were analyzed
to identify a possible historical analog to potential climate change, and
present-day, vegetation-climate classifications of plant communities were
discussed. In another approach, pollen abundances were correlated through-
regression techniques to climatic variables, such as temperature and rainfall.
Response surfaces were then constructed. Such empirical response surfaces are
available for many North American species. In general, the rate of climate
change was identified as an uncertainty that has not yet been adequately
addressed.
Another approach, still in the early stages of development, was the use
of vegetation models that represent forest dynamics by a series of stochastic
and deterministic equations. Exposure-response relationships based on
empirical data or process-oriented theory were used to model the response of
individual plant taxa. Currently these models are limited because the rate of
climate change has not yet been considered, and because only North American
communities have been modeled.
The discussion of effects of sea-level rise on wetlands emphasized
effects on coastal wetlands, including inundation, salt-water intrusion,
decline in peat formation, and shoreline erosion. Modifying factors, such as
the degree of protection of shore-front property were also discussed.
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Stratospheric Modification EPA/OAR 1987 Page C-20
Exposure Assessment
The exposure assessment component of the model is shown in the first four
modules of Figure C2-1 and, for the purposes of this summary, also in part of
the fifth module. Based on assumptions regarding production scenarios and
policy alternatives, the emissions of stratospheric modifying gases were
predicted. The atmospheric module then estimated the amount of ozone
depletion associated with the emissions. The effects module estimated lateral
ozone depletion from the one-dimensional model and estimated the amount of UV-
B increase associated with the predicted depletion. The National Academy of
Science mid-range estimate of climate sensitivity to a doubling of C02 was
used to predict temperature change. Broad changes in sea level rise by the
year 2100 were estimated for a range of warming scenarios.
Uncertainty in the exposure assessment was quantified in several ways in
the different modules. Uncertainty in estimates of the production of
stratospheric ozone-depleting gases was addressed by the use of five different
sensitivity cases, using varying assumptions regarding future populations and
economic growth. Different policy alternatives for controls on future use
were specified. The output of the atmospheric module was presented as a range
based on a given range of uncertainty for estimates of future population arid
economic growth.
Other uncertainties could not be quantitatively incorporated into the
modules. For example, possible changes in technology were not included in the
emissions module. The atmospheric module does not have the resolution to
predict regional effects. Finally, factors that are as yet undefined or
poorly understood, such as biogeochemical cycles, could not be incorporated.
Risk Characterization
The model is deterministic in that a unique output or result is produced
by a specified input; there are no stochastic elements within the model. The
exposure assessment component of the model estimated future emissions of
stratospheric modifying chemicals. The resulting depletion of ozone, increase
in UV-B, rise in equilibrium temperature, and rise in sea level were
quantitatively modeled. The effects of increased UV-B on soybean yield and
northern anchovy mortality rate were presented as ranges derived from case
studies and on-going research. Changes in forest community structure were
assessed qualitatively based on a doubled CC*2 scenario, but the rate of change
was not considered. A range of percent coastal wetland loss in U.S. was
estimated based on an analysis of topographic maps and wetlands inventories,
and on a consideration of possible protection of shorefront property.
The model has been partially validated; current measurements of ozone
were compared with model predictions. Because of the global nature of the
problem, it will be very difficult to validate the predicted effects.
However, the changes in plant communities predicted to occur with climate
change could be validated/calibrated using historical (paleobotanical) data.
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Stratospheric Modification EPA/OAR 1987 Page C-21
2.2.3 Operational Resource Requirements
The level of effort and expertise needed to develop the model was very
high. As it exists now, a moderate level of effort, expertise, and a computer
would be needed to use the model to adequately address different regulatory
and gas production scenarios.
2.2.4 Summary
There are many limitations and uncertainties in the models, exposure -
response curves, and effects data used to predict the effects of stratospheric
modification. Nevertheless, the model provides a consistent framework by
which different future scenarios and policy options can be evaluated. In
addition, the approach is flexible enough so that changes or additional
modules can be incorporated as the state of knowledge improves.
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AWQC EPA/OWRS 1986 Page C-22
3.0 METHODS/ASSESSMENTS DEVELOPED UNDER THE CLEAN WATER ACT
3.1 GUIDELINES FOR DERIVING NUMERICAL NATIONAL WATER QUALITY CRITERIA FOR
THE PROTECTION OF AQUATIC ORGANISMS AND THEIR USES (EPA/OWRS 1986)
3.1.1 Introduction
The Environmental Research Laboratories at Narragansett, Duluth, and
Corvallis developed guidelines for the Office of Water Regulations and
Standards (OWRS) for establishing limits to chemical concentrations in surface
water that are protective of aquatic organisms, human health, and some
recreational activities. The guidelines address the development of ambient
water quality criteria (AWQC) for two measurement averaging periods: 1 hour
and 4 days. These two criteria are known as the acute and chronic criteria,
respectively.
3.1.2 Description of Method
Receptor Characterization
Aquatic organisms are the primary focus of this method. The methodology
is designed to establish chemical concentration levels which, if not exceeded,
should be protective of 95% of the species in the aquatic community. The
method is designed for generic receptors, but can be modified to cover site-
specific or locally-important species (e.g., commercially important species).
Toxicity data from tests on the most sensitive life stage are recommended, but
are not essential. Niche characteristics are not directly considered in the
method unless field data for bioaccumulation of the chemical are available.
The method is more fully developed for animals than for plants primarily
because the test methodology is more refined for animal species.
Hazard Assessment
The methodology is designed to establish chemical concentrations which,
if not exceeded more than once in three years, will be protective of 95% of
aquatic species. The necessary data include acute and chronic toxicity data
for animals and plants (including algae), bioconcentration or bioaccumulation
factors, and either FDA tissue action levels or chronic wildlife feeding/field
study results. The most rigorously defined criterion is the 1-hr maximum
concentration. Data requirements for calculation of this criterion include
toxicity data for animal species from a specified taxonomic range of orders
and families. The maximum continuous concentration (4-day) value is taken as
the lowest of either an animal chronic value, a plant toxicity value, or a
tissue residue value.
The chronic toxicity value for aquatic animals can be calculated from
chronic bioassay results in the same manner as the 1-hour maximum
concentration, that is, by using toxicity data for animal species from a
specified taxonomic range of orders and families. An alternative method is Co
use a chemical-specific acute:chronic toxicity ratio applied to the acute
value derived above. The ratio method requires considerably fewer daca Chan
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AWQC EPA/OWRS 1986 Page C-23
the chronic bioassay data set requirements. A toxicity value for plants is
simply taken as the lowest reported toxic concentration for a plant or algal
species. A specific plant toxicity endpoint is not specified. A tissue
residue value is calculated from data on either FDA action levels or chronic
wildlife feeding or field studies and bioconcentration or bioaccumulation
factors.
Guidance is provided on appropriate test methods, test species, and life
stages to be included in the data sets. Also, a regression method is included
for calculation of acute and chronic toxicity levels when the toxicity of the
chemical is affected by background water quality (e.g., water hardness). Salt
and fresh water criteria are calculated separately. If a criterion is above a
documented toxic threshold for a commercially or recreationally important
species, the criterion can be replaced with the lower toxic level. The method
is designed to calculate a criterion .that corresponds to a value for a species
representative of the lowest fifth p'ercentile of species ranked according to
sensitivity, and so is designed to be protective of 95% of the tested species.
Guidelines are set for the development of single chemical criteria only;
mixtures are not addressed.
Exposure Assessment
Exposure to aquatic organisms is assumed to occur only through water,
although exposure to predators through the food chain is assessed by the
residue value. The estimation or measurement of exposure in the environment
is not a component of criteria development.
Risk Characterization
An assessment of the potential for adverse ecological effects can be
conducted by comparing the criteria generated by the above method to
concentrations of chemicals measured or estimated in the environment, as
described in Sections 5.2 and 7.1.1. Criteria are not calculated for
chemicals with insufficient taxonomic breadth in the available data. That the
method is protective of 95 percent of aquatic species and that this level of
protection prevents adverse ecosystem-level effects could be validated, but
only by extensive field monitoring projects.
3.1.3 Operational Resource Requirements
A computer program that can be run on a personal computer is available
for performing the calculations. Large toxicity data bases exist for plant
and animal species which can be used by this method. However, the range of
required species which have been tested with each chemical restricts the
number of chemicals for which an acute and chronic criterion can be
calculated. The level of effort is low, particularly if the computer program
is available. However, validation of literature reports requires a high level
of expertise.
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AWQC EPA/OWRS 1986 Page C-24
3.1.4 gimrpju-y
The guidelines were developed to calculate water quality criteria that
will be protective of populations of most aquatic organisms. The guidelines
are not concerned with what effects, if any, will occur in the ecosystem if
concentrations reach or exceed the benchmarks. Criteria are developed for
both acute and chronic exposure scenarios. The acute criterion is a
prediction of a concentration threshold above which acute toxicity to some
species might occur. It is derived primarily from laboratory acute toxicity
data. The chronic benchmark can be derived from one of three data sets. No
weighting is given to the quality or quantity of data within each data set.
The lowest of the three chronic values (i.e., animal, plant, and tissue
residue values) is chosen arbitrarily, even though the calculation schemes are
different. The calculation schemes themselves are straightforward. However,
validation of input data requires a high level of expertise in the area of
aquatic toxicology.
Strengths of the approach are its wide applicability and rigor in
requiring a minimal set of data to calculate a criterion. Limitations are
that community or ecosystem effects are not directly assessed and that
guidelines have not yet been established for chemical mixtures.
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Approach to Exposure and Risk EPA/OWRS 1983 Page C-25
3.2 AN APPROACH TO ASSESSING EXPOSURE TO AND RISK OF ENVIRONMENTAL
POLLUTANTS (EPA/OWRS 1983)
3.2.1 Introduction
A.D. Little, Inc. prepared this document for the Office of Water
Regulations and Standards (OWRS) to provide guidelines for conducting risk
assessments for waterborne pollutants and to identify priorities for further
development. Part of the guidelines presented are for assessment of risks to
non-human biota. A schematic diagram of the proposed approach to risk
assessment is shown in Figure C3-1. The approach is qualitative. The result
of the assessment is a listing or summary of species, locations, exposures,
and effects levels, indicating the combinations most likely to result in high
risks.
3.2.2 Description of Method
Receptor Characterization
Emphasis is placed on aquatic organisms, although the procedure is
intended to be used to assess risks to any organisms coming in contact with
water (e.g., waterfowl). The guidelines suggest that the assessor identify'"
specific communities or species exposed or potentially exposed to the
pollutant. Representative sensitive species should be identified and their
ranges delineated. In addition, behavior patterns (i.e., migration, age
structure) that might increase or decrease the potential for exposure should
be described.
Hazard Assessment
The proposed hazard assessment is based on a literature review. Because
the procedures are intended to be used for the priority pollutants, it is
assumed that Ambient Water Quality Criteria will be available. The result of
the hazard assessment is a tabulation of effects by species affected (e.g.,
fish, aquatic invertebrates) and by type of effect (e.g., lethal and
sublethal). No methods for extrapolation between species or time frames are
suggested. When available, the relationship between toxicity modifying
factors and effects should be included in the summary.
Exposure Assessment
Potential for exposure is estimated using either environmental fate and
transport models or field monitoring. Exposure analysis components include
identification of the concentration of the chemical, the areal extent of
contamination, and temporal variation in levels of contamination.
Risk Characterization
It is assumed that incomplete data will necessitate qualitative
environmental assessments the majority of the time. An initial assessment is
made to determine whether the chemical is likely to have (1) no effect on
-------
Approach to Exposure and Risk
EPA/OWRS 1983
Page C-26
FIGURE C3-1
SCHEMATIC DIAGRAM OF EPA/OWRS (1983)
APPROACH TO RISK ASSESSMENT
I KrH.I'S ANALYSIS
KXri>S11RK ANALYSIS
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-------
Approach to Exposure and Risk EPA/OWRS 1983 Page C-27
aquatic organisms; (2) some effects on certain sensitive species; (3) effects
on most species; or (4) effects on all species. Exposure and hazard
information is then examined to refine the rough analysis. Possible
refinements include comparing time-dependent patterns of contamination levels
(e.g., persistence, seasonal fluctuations) with species activity patterns and
considering toxicity modifying factors such as hardness. Expected results can
be validated with information on fish kills or field sampling programs. The
goal of the assessment is to establish "key intersections" between exposure
and effects data with regard to specific species and geographic locations.
The product of the analysis is a listing or summary of species, locations,
exposures, and effects levels, with an indication of the combinations most
likely to result in high risks.
3.2.3 Operational Resource Requirements
The approaches proposed in this.document can be made more sophisticated
or simple, depending on needs and resources. If collection of field data is
necessary, resource needs will be high. Although the approach is simple and
qualitative, it requires a high level of judgment and knowledge of aquatic
systems.
3.2.4 Summary :
The guidelines presented in this document offer a systematic way to
consider available data and design needs for possible field work. The
strengths of the approach are that it is very flexible and relies on common
sense and professional judgment. The limitations of the approach are that the
results are not quantified. In addition, little guidance is provided on the
approaches for extrapolating from available data to species and effects of
concern, and the specific criteria upon which to base conclusions (e.g., NOEL,
LC50) are not defined.
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Water Quality-based Permitting EPA/OWRS 1985, 1987 Page C-28
3.3 PERMIT WRITER'S GUIDE TO WATER QUALITY-BASED PERMITTING FOR TOXIC
POLLUTANTS / TECHNICAL SUPPORT DOCUMENT FOR WATER QUALITY-BASED TOXICS
CONTROL
3.3.1 Introduction
Both the Clean Water Act and promulgated Federal regulations require that
all National Pollutant Discharge Elimination System (NPDES) permits include
limitations to achieve all applicable state water quality standards. All
states have numerical standards for some individual toxicants, as well as
narrative standards for pollutants. One of the most common narrative
standards is that a state's waters shall be free from substances in
concentrations or combinations toxic to humans, wildlife, or aquatic life.
This standard can be used to limit both individual toxicants and whole
effluent toxicity. The Permit Writer's Guide to Water Quality-Based
Permitting for Toxic Pollutants (EPA/OWRS 1987) provides procedural
recommendations to state and Federal NPDES permit writers. The Technical
Support Document (TSD) for Water Quality-Based Toxics Control (EPA/OWRS 1985)
provides additional guidance and explanation of the two basic approaches to.
water quality-based toxics control: the whole-effluent approach and the
chemical-specific approach. Either or both approaches can be used for setting
effluent limits or monitoring compliance. This review describes only the
approaches recommended to ensure that state waters are free from substances
toxic to aquatic life.
The whole-effluent approach to toxics control involves toxicity testing
of the effluent. The measure of whole-effluent toxicity, information on
mixing zones, and design flow of the receiving waters are then used to
estimate a wasteload allocation/total maximum daily load (WLA/TMDL) for
effluent discharge. The chemical-specific approach involves use of a water
quality criterion (e.g., state numerical water quality criteria which are
often derived from EPA ambient water quality criteria) and exposure modeling
to establish an effluent discharge limit.
3.3.2 Description of Method
Receptor Characterization
The initial step in the water quality-based approach to toxic effluent
control is determining the level of water quality that must be maintained to
allow continuation of the state-designated use of the water body (e.g., cold
water fishery). Further receptor characterization depends on whether the
whole-effluent approach or the chemical-specific approach is employed. If the
whole-effluent approach is followed, the TSD recommends use of three different
test species so that a species sensitive to the toxicants in the effluent is
likely to be tested. If the chemical-specific approach is followed, the
receptor characterization is that associated with the development of the
numerical criteria (see review of EPA/OWRS 1986 Ambient Water Quality Criteria
methodology).
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Water Quality-based Permitting EPA/OWRS 1985, 1987 Page C-29
For the whole-effluent approach, it is not possible to predict a. priori
which species is likely to be most sensitive because different species exhibit
different sensitivities to toxicants (e.g., trout are very sensitive to oxygen
depletion but are relatively insensitive to certain toxicants). An analysis
of species sensitivity ranges (as identified in the EPA Ambient Water Quality •
Criteria documents) indicated that if tests are conducted on three particular
species (Daphnia magna. Pimephales oromelas. and Lepomis macrochirus), the
most sensitive of the three will exhibit an LC5Q value within one order of
magnitude of the most sensitive of all species tested. If fewer than three
species are tested, a safety factor of 10 should be applied to the lowest
observed LC5Q value to extrapolate to more sensitive species. The TSO
recommends against using resident species for the whole-effluent toxicity
tests unless it is required by state statute or some other binding factor.
The reason is that testing of resident species is more costly, more difficult,
and subject Co more variability than testing of standard laboratory species.
Hazard Assessment
The hazard assessment of the water quality-based approach to toxics
control depends on whether the whole-effluent approach or the chemical-
specific approach is followed. For the whole-effluent approach, toxicity .
tests are conducted on three test species using a series of dilutions of the
whole-effluent. For the chemical-specific approach, the hazard assessment has
already been conducted in the establishment of numerical ambient water quality
criteria (see review of EPA/OWRS 1986 Ambient Water Quality Criteria
methodology).
For the whole-effluent approach, an effluent sample is collected and
diluted in test chambers; the dilutions are usually 100%, 30%, 10%, 3%, 1%,
and a control. The receiving water is frequently used to dilute the effluent
because it more closely simulates the effluent/receiving water interactions.
The required measures of effluent toxicity are the LCjQ (the effluent
concentration expressed as the dilution at which 50% of the test organisms are
killed) and the no-observed-effect-level (NOEL; the highest effluent
concentration at which no unacceptable effect will occur even following
continuous exposure). Because of the inverse relationship between toxicity
and the reference concentration (e.g., the lower the LC5Q or the NOEL, the
higher the toxicity of the effluent), concentration-based toxicity
measurements are translated into acute (TUa) or chronic (TUc) "toxic units."
The toxic unit is simply 100 divided by the toxicity measure:
TUa - 100 TUc - 100
LC50 NOEL
where the LC5Q or NOEL is expressed as percent effluent in the dilution water.
Thus, an effluent for which the 10% dilution killed 50 percent of the test
organisms is an effluent containing 10 TUa. An effluent contains 20 TUc if
the highest concentration that did not produce adverse effects over a long
exposure period was a 5% dilution.
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Water Quality-based Permitting EPA/OWRS 1985, 1987 Page C-30
Exposure Assessment ^
Some States have numeric criteria for whole effluent toxicity, often
stated as end-of-the-pipe acute toxicity limits, that do not depend on an
exposure assessment. Otherwise, effluent concentrations in both the mixing
zone(s) surrounding the effluent discharge and in the remainder of the
receiving water body must be estimated on the basis of the design flow of the
receiving waters. Design flow is a hydrological condition which describes a
low flow of a flowing water body and is calculated from the historical flow
record. The level of effluent control is to be set on a "worst case" exposure
scenario to protect aquatic life during low flow conditions, when the dilution
potential of the receiving water body is lowest. Design flows can be
calculated using an EPA recommended computer simulation program (DFLOW) or
other methods. The goal of the exposure assessment is to determine the lowest
flow expected for a specified period of time with a specified frequency of
occurrence for comparison with the reference toxic units. The lowest flow
averaged over any consecutive 7 day period that is expected to occur in a
typical 10 year period (7Q10) is the recommended design flow for calculating
dilution for the chronic ambient water quality criterion or for the whole-
effluent identified NOEL. The lowest flow averaged over any 1 day period that
is expected to occur in a typical 10 year period (1Q10) is the recommended .
design flow for calculating dilution for the acute criterion or the whole-"
effluent LC5Q value.
Risk Characterization
Many state standards allow a zone of mixing around the effluent discharge
point in which less stringent ambient water quality criteria apply than in the
remainder of the receiving water. EPA's policy on mixing zones, described in
the 1983 Water Quality Standards Handbook, is that any mixing zone should oe
free from materials which cause acute toxicity to aquatic life. Acute
toxicity is of concern for organisms passing through the mixing zone on a
daily or seasonal basis. Both acute and chronic toxicity are of concern for
the remainder of the receiving water body.
Tiered testing, similar to that used under the Toxic Substances Control
Act (TSCA) when testing a new chemical product for potential hazard, is
recommended to provide a cost-effective method of obtaining the necessary data
for toxic effluent control. At the screening level, the goal is to separate
situations in which impacts on aquatic systems are improbable from those in
which impacts are possible or probable. Screening level considerations
include the dilution potential of the receiving water, the type of industry,
the type and volume of industrial input, any existing data on toxic
constituents in the effluent, and any history of toxic impact on receiving
waters or compliance problems. If, during the screening analysis, a case is
identified wherein impacts on aquatic organisms are possible or likely,
definitive data gathering is then required using either the whole-effluent
approach or the chemical-specific approach, or both.
In the whole-effluent approach, the recommended acute criterion for the
acute design flow (1Q10) is 0.3 TUa. The adjustment factor of approximately
one-third is used to extrapolate from a 50 percent mortality level (LC50) to a
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Water Quality-based Permitting EPA/OWRS 1985, 1987 Page C-31
1 percent mortality level (LC^). The criterion can be applied at different
locations in the receiving water depending on the dilution situation. The
recommended chronic criterion for the chronic design flow (7Q10) is 1 TUc.
The waste load allocation (WLA) requirement is calculated on the basis of the
toxicity criterion and the dilution factor based on the receiving water design
flow as well as the effluent design flow. A separate WLA is calculated for
the acute and chronic exposure scenarios, and an effluent performance level
that will meet each WLA requirement is back-calculated. Permit limits are
derived from the more restrictive performance level (the acute or the chronic
design flow).
3.3.3 Operational Resource Requirements
The effort and cost of following the recommended analysis and testing
procedures for NPDES permitting depends on the results of the dilution
screening analysis. The TDS identif-ies five levels of dilution potential in
the receiving water body and recommends increasing numbers of toxicity tests
as the dilution potential of the receiving water declines. Potentially, both
acute and short-term chronic toxicity tests might be required for three
species. Compliance monitoring might also require effluent toxicity testing
on a monthly basis with the one species determined to be most sensitive to. Che
effluent.
3.3.4 Summary
The Water Quality-Based Toxics Control methodology includes two
approaches to defining limits to toxic effluent discharges into surface
waters. The whole-effluent approach involves determining the toxicity of the
effluent mixture using bioassays with three species of aquatic organisms. The
chemical-specific method involves the application of state water quality
criteria for the designated uses of the receiving water. Each approach has-
advantages and disadvantages and is complementary to the other. EPA
recommends using both approaches, as appropriate, to a given permitting
situation.
The principal advantages of the whole-effluent approach compared with the
chemical-specific approach are (1) the combined toxicity of all constituents
in a complex effluent is measured, (2) the bioavailability of the toxic
constituents is assessed, and (3) the effects of synergistic or antagonistic
interactions of the effluent constituents are accounted for.
The principal disadvantages of the whole-effluent approach compared with
the chemical-specific approach are (1) treatment systems are more easily
designed to meet chemical-specific requirements, (2) engineers and permit
writers are more familiar with chemical-specific approaches, (3) properties of
specific chemicals in complex effluents (e.g., bioaccumulation) are not
assessed.
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Biological Criteria Ohio EPA 1987a, 1987b, 1988 Page C-32
3.4 BIOLOGICAL CRITERIA FOR THE PROTECTION OF AQUATIC LIFE
3.4.1 Introduction
Since 1980, Ohio EPA has been conducting biological evaluations of state
surface water quality to quantify the attainment or non-attainment of state
designated surface water uses. Field measurements of the fish and
macroinvertebrate sub-communities are used to calculate three indices of
biological integrity: (1) Index of Biotic Integrity (IBI, based on the fish
sub-community), (2) Modified Index of Well-Being (Iwb, also based on fish),
and (3) Invertebrate Community Index (ICI, based on the macroinvertebrate sub-
community). Ohio EPA uses the three indices in conjunction with chemical-
specific ambient water quality criteria and toxicity testing for discharge
permitting to improve and maintain the biological integrity of the state's
surface waters (Ohio EPA 1987a, 1987b. 1988).
Ohio EPA has found the indices to meet several important criteria. The
indices depend on actual biological measures, rather than surrogate measures
such as contaminant concentration. The measures indicate effects at several
trophic levels. The response ranges of the indices are suitable for the
regulatory needs. The indices are sensitive to the environmental conditions
of interest. The indices are reproducible and precise within defined and
acceptable limits. Finally, the signal-to-noise ratio for the combination of
the three indices is high.
3.4.2 Description of Method
Receptor Characterization
The Ohio EPA definition of biological integrity depends on the biological
conditions exhibited by the "least impacted" surface water habitats, rather
than "pristine" habitats which would represent unattainable conditions over
most of the state. Ohio EPA uses an ecoregion approach to identify reference
aquatic communities, relying on Omernik's (1987) classification of aquatic.
ecoregions of the United States from maps of land-surface form, soils,
potential natural vegetation, and land use. Within each of the five
ecoregions, three types of rivers/stream habitats are defined based on size
and water flow: wading sites, boat sites, and headwater sites.
The fish and macroinvertebrate sub-communities were selected for
evaluation of surface water quality because there is sufficient information
concerning their life-histories, distribution, and tolerances and also because
these sub-communities are dependent upon the other sub-communities (e.g.,
plants, microinvertebrates) for their well-being. The fish sub-community is
the most conspicuous and is generally of concern for its commercial or sport
value.
The Index of Biotic Integrity (IBI) incorporates 12 different metrics of
the fish sub-community to derive a final IBI score. The value of each
individual metric measured at an evaluation site is compared with the value
expected on the basis of the analysis of "least impacted" reference sites in
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Biological Criteria Ohio EPA 1987a, 1987b, 1988 Page C-33
the same ecoregion with the same water flow characteristics. Ratings of 5, 3,
or 1 are assigned to each of the 12 metrics based on whether the evaluation
site value approximates, deviates somewhat from, or deviates strongly from the
expected value, respectively. Expected values are derived from the data
collected at the "least impacted" reference sites. The 12 community metrics
were modified from those proposed by Karr (1981) to include measurements that
could be reliably made given the sampling gear required for each stream size.
The IBI metrics include the total number of species, three metrics related to
the total number of species of a specified type of fish (e.g., sucker, minnow,
sunfish), the number sensitive (i.e., intolerant) species, the percent
tolerant species, the percent omnivorous species, the percent insectivorous
species, the percent top carnivore or pioneering species, the total number of
individuals, the percent hybrids, and the percent diseased individuals or fish
exhibiting deformities, eroded fins, lesions, or external tumors.
The Index of Well-Being (Iwb) for the fish sub-community incorporates
four indices: numbers of individuals, biomass, the Shannon diversity index
based on number, and the Shannon diversity index based on weight. Number and
biomass data are obtained from pulsed D.C. electrofishing catches, where
sampling effort is normalized on the basis of distance. Ohio EPA recently
developed a modification of the Iwb that makes the index more sensitive to ,
environmental disturbances. In the modified Iwb, any of 13 highly tolerant
species, hybrids, or exotic species are eliminated from the numbers and
biomass components of the Iwb but still included in the two Shannon diversity
indices. Ratings are assigned to each of the four indices on the basis of
similarity to the reference sites.
The Invertebrate Community Index (ICI) is an index of the
macroinvertebrate sub-community as measured on artificial substrates
introduced into the surface water for a specified period of time. The ICI is
a modification of the IBI for fish and consists of 10 metrics, including the
total number of taxa, number of mayfly taxa, number of caddisfly taxa, number
of dipteran taxa, percent mayflies, percent caddisflies, percent midges (tribe
tanytarsini), and percent tolerant organisms. Mayflies, caddisflies, and
tanytarsinid midges are important components of undisturbed stream
macroinvertebrate sub-communities that range from highly to moderately
sensitive to pollutant stress. Ratings are assigned to each of the 10 metrics
on the basis of similarity to the reference sites.
Regional criteria for the IBI, modified, Iwb, and ICI were established on
the basis of the measurements of each metric from the reference sites.
Ecoregional criteria for the warmwater habitat (WH) use designation were
established as the 25th percentile value from the reference site values for
each ecoregion. The exceptional warmwater habitat (EWH) use designation
criteria were set at the 75th percentile value. The 25th percentile for the
WWH use designation was chosen to compensate for the inclusion of marginal
sites in the original reference site data base. Ecoregional criteria for
coldwater habitats, however, have not yet been developed.
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Biological Criteria Ohio EPA 1987a, 1987b, 1988 Page C-34
Hazard Assessment
A hazard assessment is not part of the biological criteria methodology.
However, the biological criteria are intended for use in conjunction with
chemical identification, chemical criteria, and toxicity testing to assist
water quality management decisions.
Exposure Assessment
Exposure characterization is not part of the biological criteria
methodology. The fish and macroinvertebrate sub-communities represent
relatively long-lived organisms that have integrated the effects of continuous
exposure to the pollutant and other stressors over a period of years.
Risk Characterization
The biological indices from an evaluation site can be used in several
ways. The indices can be compared with the criteria values for the ecoregion
to establish the attainment or non-attainment of a designated use. The
indices for upstream and downstream locations can be compared to help identify
a hazardous discharge. Indices from the same site can be'followed over time
to measure the effectiveness of a regulatory control. For some types of . :
pollutants, the ICI measure of the macroinvertebrate sub-community is a more
sensitive indicator than either fish index, and for other types of stress, the
fish sub-community indices can be more sensitive. All three indices are
therefore used to indicate the quality of Ohio surface waters. Figure C3-2
illustrates how the three indices have been used to illustrate the general
improvement in water quality over the last 10 years at selected sampling
stations in the state of Ohio. Figure C3-3 illustrates how the indices can be
used to document environmental degradation associated with one or more point
sources of pollutant stress with distance along a river.
3.4.3 Operational Resource Requirements
Ohio EPA provided a cost comparison of fish sub-community and
macroinvertebrate sub-community evaluations of water quality with chemical and
physical sampling of surface water and acute and acute/chronic bioassay tests
on effluents of the type used for National Pollutant Discharge Elimination
system (NPDES) permitting. While chemical/physical sampling (4-6 samples per
site) costs between $1,500 and $1,700, and bioassays can run from $3,200 for a
screening level assay to $8,000 - $12,000 for a 7-day bioassay, a -
macroinvertebrate evaluation typically costs $700 and fish community sampling
costs Ohio EPA between $670 and $900 (2 to 3 passes per site). Thus, the
biological evaluation methodology is cost competitive with the more commonly
used measures of surface water quality. The initial development of Biological
Criteria, however, required a substantial effort in sampling a large number of
"least impacted" reference sites.
-------
Biological Criteria
Ohio EPA 1987a, 1987b, 1988
FIGURE C3-2
Page C-35
BIOLOGICAL IHDICES OF SURFACE WATER OJUALITY IN THE SCIOTO RIVER
(Source: Ohio EPA 1988)
H
Q
H
O
UJ
H
u.
H
O
O
z
r..,,.,,.l....l,.,.l,,.,l,,,.,....,....,....,,...,,,.I
Y E A
Fieure C3-2 Results of the Invertebrate Community Index (ICI), Index of
Biotic Integrity (IBI), and Modified Index of Well-being (Iwb) at selected
!amp ing locations in the middle Scioto River between 1974 and 1987. Shading
indicates boundaries between exceptional (EWH), good (WWH). fair, poor, and
very poor conditions and the variability of each index. Sampling was
conducted during July-October of each year.
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Biological Criteria
Ohio EPA 1987a, 1987b, 1988
FIGURE C3-3
Page C-36
LONGITUDINAL TRENDS IN BIOLOGICAL INDICES OF SURFACE WATER QUALITY
IN THE CUYAHOGA RIVER
(Source: Ohio EPA 1988)
1" 'f I ')' " I" " M 1" i""
RIVER MIl_E
Figure C3-3. Longitudinal trend of the Invertebrate Community Index (1984,
1986) and Modified Index of Weil-Being (fish; 1984, 1985, and 1986) in the
Cuyahoga River study area. Shading indicates boundaries between exceptional
(EWH), good (WWH), fair, poor, and very poor conditions and the variability of
each index. Sampling was conducted during June-September of each year.
Horizontal arrow indicates direction of flow; environmental influences are
indicated at the top.
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Biological Criteria Ohio EPA 1987a, 1987b, 1988 Page C-37
3.4.4
Ohio EPA has developed and is using three indices of biological integrity
to measure attainment of the goals of the Water Quality Act (WQA) . Unlike
chemical criteria which serve as surrogate measures of the attainment of the
biological goals of the WQA or effluent toxicity testing to control
chemicaldischarges , the indices incorporate direct measurements of the
structure of biological communities, representing a top-down approach. There
are several advantages to using measurements of the fish and invertebrate
communities to assess water quality. One advantage is that the organisms have
been exposed continuously to the actual conditions and history of pollutant
stresses in the receiving waters. Extrapolation between responses of
organisms in the laboratory and responses of organisms in the field are not
required. In addition, stressors other than toxic substances for which
laboratory data are available can be- detected. The organisms integrate
exposure over time and space, and the condition of the resident community is
the result of the full history of environmental conditions, including both
common and extreme events .
Some difficulties that have discouraged the development of biological.-.
criteria and community- level indices in the past have been overcome through
standardization of field sampling methods, acceptance of several types of
measures as indicative of biological integrity, calibration of indices on the
basis of ecoregion and surface water body type, and measurement of the indices
for a large number of reference sites. Over the past 10 years, Ohio EPA has
found that the signal-to-noise ratio and reproducibility of the indices are
sufficiently high to provide useful indications of changes in the biological
community in response to pollutants or other stresses.
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Biological Criteria Ohio EPA 1987a, 1987b, 1988 Page C-38
3.4.5 References
Karr, J.R. 1981. Assessment of biotic integrity using fish communities.
Fisheries 6(6):21-27.
Larsen, D., Omernik, J.M., Hughes, R.M., et al. 1986. Correspondence between
spatial patterns in fish assemblages in Ohio streams and aquatic
ecoregions. Env. Mgmt. 10(6):815-828.
Omernik, J.M. 1987. Ecoregions of the conterminous United States. Ann.
Assoc. Amer. Geogr. 77(1):118-125.
Whittier, T.R., Larsen, D.P., Hughes, R.M., et al. 1987. The Ohio Stream
Regionalization Project: a Compendium of Results. US Environmental
Protection Agency, Freshwater Research Laboratory, Corvallis, OR.
EPA/600/3-87/025. 163 pp.
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Fish Flesh Criteria NYS/DEC 1987 Page C-39
3.5 NIAGARA RIVER BIOTA CONTAMINATION PROJECT: FISH FLESH CRITERIA FOR
PISCIVOROUS WILDLIFE (NYS/DEC 1987)
3.5.1 Introduction
To interpret the results of the Niagara River fish tissue monitoring
program, the Niagara River Toxics Committee (NRTC 1984) of New York State
recommended that residue criteria be established, if none were available, for
the chemicals that had been detected. The two primary objectives of this
report were (1) to develop fish flesh criteria for 19 organochlorine chemicals
that will protect piscivorous wildlife, and (2) to evaluate a methodology for
deriving such criteria where toxicology data are.unavailable for wildlife
species of concern. In the past, fish flesh criteria for the protection of
wildlife generally have been derived from wildlife feeding studies, of which
few are available. The NYS/DEC approach is to use the extensive laboratory
animal toxicology data base employed'for the derivation of human health
criteria to extrapolate to wildlife species. In the following sections, we
review this methodology for deriving fish flesh residue criteria.
3.5.2 Description of Method
Receptor Characterization
A list of 18 species of piscivorous wildlife which are current or former
inhabitants of the Niagara River valley were selected as the ecological
targets of concern. For each of these species, body weight, daily food
consumption by weight, and food habits were determined. From these data, the
mammal and bird with the greatest ratios of daily food consumption to body
weight were selected for use in calculation of fish flesh criteria. Mink,
with an average body weight of 1 kg and daily food consumption of 0.15 kg/day,
were, selected as the representative mammalian target. Because several species
of birds consume an amount of food equal to about 20 percent of their body
weight daily, a "generic" bird was selected with a weight of 1 kg and a daily
food consumption of 0.2 kg.
Hazard Assessment
The results of laboratory animal feeding experiments are extrapolated to
fish flesh criteria for wildlife using the following general formula:
(Toxicitv Value (mg/kg-dav) x Uncertainty Factor x Target Species Weight (kg)1
Target Species Daily Intake (kg/day)
The Toxicity Value is either a no-observed-effect-level (NOEL), lowest-
observed-effect-level (LOEL), or cancer risk-based dose corresponding to a
1/100 risk of cancer. Where a chronic NOEL for a sensitive species is
unavailable, the uncertainty factor is applied as follows:
0.1 is used to estimate a chronic NOEL from subacute data;
0.2 is used to estimate a chronic NOEL from a chronic LOEL; and
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Fish Flesh Criteria NYS/DEC 1987 Page C-40
0.1 represents the interspecies uncertainty factor to use when chronic
data are available from only one or two species in the same class.
Only laboratory data from mammalian species are used to extrapolate to a
mammalian wildlife species (i.e., mink), and only laboratory data from birds
are used to extrapolate to a wildlife bird species (i.e., the "generic" bird).
Toxicity endpoints of concern included mortality, body weight gain, liver
and kidney weight (as a percentage of body weight), liver and kidney
micropathology, and reproductive losses. The interspecies uncertainty factor
is based on a literature review that supports a 10-fold or more range in
sensitivity of species to thoroughly tested organochlorines. The factor of
0.1 was selected to estimate a chronic NOEL from subacute data on the basis of
a review by Weil and McCollister (1963) in which the authors found that a 10-
fold factor would cover 95 percent of. the chemicals tested for short-term
(e.g., 30 to 90 days) versus long-term (e.g., two years or lifetime) exposure.
Weil and McCollister (1963) also presented data that justified using a factor
of 0.2 to extrapolate from a LOEL to a NOEL (all ratios of LOEL/NOEL were 10
or less, and 92 percent were 5 or less). The NYS/DEC also recommends using
supplementary data when available, such as epidemiological field studies.
As one method of validating this approach, NYS/DEC compared the fish
flesh residue criteria developed for five chemicals, using the method outlined
above including the uncertainty factors, with toxicity data for the actual
target wildlife species of concern. For PCBs, DDT, aldrin/dieldrin, and
mirex, but not endrin, at least one of the lab species-based non-carcinogenic
criteria was lower than target species criteria. In the case of endrin, the
mallard exhibit 3. very high sensitivity to the chemical. Thus, using
laboratory animal data in conjuction with uncertainty factors should result ir
fish residue levels that would usually be protective of wildlife species.
This method does not, however, address the possible additive or synergistic-
effects of mixtures of similar compounds or the susceptibility of wildlife in
naturg to toxic substances.
Exposure Assessment
The most sensitive bird (i.e., "generic") and mammal (i.e., mink) species
are assumed to consume a diet consisting of 100 percent fish. The NYS/DEC
points out that biomagnification of contaminants might continue in the
terrestrial food chain, such as when eagles consume gulls that have consumed
contaminated fish. The method for deriving fish flesh criteria outlined above
is not protective of terrestrial animals in higher trophic positions than the
piscivorous target wildlife species.
Risk Characterization
To assess risk to wildlife fish consumers near the Niagara River, the
fish flesh criteria were compared to residues in Niagara River fish (e.g.,
spottail shiners, alewives, smelt, coho salmon) by the quotient method (see
Review 7.1.1). Extrapolations of residue levels in spottail shiners (which
were the only species sampled on a regular basis over the years) to residue
levels in other fish species were accomplished by assuming that all of the
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Fish Flesh Criteria NYS/DEC 1987 Page C-41
organochlorine compounds were sequestered in the lipid portion of the fish,
and using a relative lipid content correction factor. The selection of a
1/100 cancer risk level as an acceptable cancer risk for wildlife was a
preliminary decision and will be studied further.
3.5.3 Operational Resource Requirements
The methodology for setting fish flesh residue criteria is simple and
easy to follow. Large toxicity data bases exist for mammalian laboratory
species because these data are used to assess human toxicity. Fewer data are
available for laboratory bird species for compounds other than pesticides.
3.5.4 Summary
The guidelines developed to calculate fish flesh residue criteria allow
extrapolation from a large laboratory animal toxicity data base to fill a
large data gap in available wildlife long-term feeding studies. The
guidelines might not be sufficiently protective of wildlife species that
consume piscivorous wildlife species, however. The guidelines are not
concerned with what effects, if any, will occur in the ecosystem if
concentrations reach or exceed the benchmarks. Nor does the risk assessment
portion address the problem of exposure to multiple contaminants. Criteria
are developed only for chronic dietary exposure scenarios. The calculation
schemes themselves are straightforward. However, validation of input data
requires a high level of expertise in the area of aquatic toxicology.
Strengths of the approach are its wide applicability and apparent
validity. Limitations are that community or ecosystem effects are not
directly assessed, that guidelines have not yet been established for chemical
mixtures, and that validation efforts have thus far been limited.
Fish residue criteria developed using this methodology could be used in
conjunction with bioaccumulation factors to adjust chronic water qualtity
criteria for substances with a strong tendency to bioconcentrate.
3.5.5 References
Weil, C., and McCollister, D. 1963. Relationship between short and long-term
feeding studies in designing an effective toxicology test. Agric. Food
Chem. 11:486-491.
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SEP for Ecological Risk EPA/OPP 1986 Page C-42
4.0 METHODS/ASSESSMENTS DEVELOPED UNDER FIFRA
4.1 STANDARD EVALUATION PROCEDURE FOR ECOLOGICAL RISK ASSESSMENT (EPA/OPP
1986)
4.1.1 Introduction
The Office of Pesticide Programs (OPP) has developed a Standard
Evaluation Procedure for carrying out ecological risk assessments to evaluate
environmental toxicology and effects data submitted in support of pesticide
registration. The approach is a modified quotient method in which estimated
environmental concentrations (EECs) are compared to environmental toxicity
endpoints; the resulting ratios are evaluated according to regulatory risk
criteria (RRCs) established by regulation or, in the case of endangered
species, by a Memorandum of Understanding between EPA and DOI.
OPP developed this approach because of several features unique to the
ecological risk evaluation of pesticides. Since pesticides are used by
deliberate and broadscale introduction into the environment, and since
pesticides are by definition toxic to at least some component: of the
biosphere, a basic set of environmental test data is required for all
pesticides before use is allowed. The approach is designed to identify
pesticides with high potential for ecological impacts, based on acute
toxicity, which would then require further testing and refinement of the risk
assessment. Finally, separate regulatory risk criteria are used to make
regulatory decisions when endangered species are the receptors of concern.
4.1.2 Description of Method
Receptor Characterization
Both aquatic and terrestrial receptors are considered under this method.
Assessments are focused at the population level, although the value of
individual members of endangered species is considered by using a more
stringent RRC. Life habit characteristics are considered when characterizing
potential receptors. For example, potential aquatic receptors are identified
and differentiated based upon the habitat in which they exist (cold or warm
water, salt or fresh water). Terrestrial species are identified and
characterized based upon habitat and dietary characteristics. Indicator
species thaC most closely represent the characteristics of potential non-
target receptors are selected for the risk assessment. Seasonal or life-cycle
characteristics and the geographic location of the potential receptors are not
considered.
Hazard Assessment
Hazard assessments are conducted at the population level, and acute and
chronic toxicity are the endpoints considered; dose-response data and
community and ecosystem responses are not evaluated. In addition, the
influence of potential toxicity modifying factors (e.g., pH, hardness,
behavioral modifications) is not considered.
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SEP for Ecological Risk EPA/OPP 1986 Page C-43
Toxicity cests that may be required for assessment are shown in Table
C4-1. The tests are arranged in a tiered system which progresses from basic
laboratory tests to applied field tests. In general, the first tier consists
of tests to determine acute endpoints (e.g., LC5QS) using common laboratory
organisms; the second tier consists of tests for sublethal effects or tests on
specific organisms; and the third tier includes field tests. The results of
each tier of tests are evaluated to determine the potential of the pesticide
to cause adverse effects and to determine whether further testing is needed.
Exposure Assessment
The exposure assessment is also conducted in a tiered manner. In the
lower tiers, environmental exposure concentrations (EECs) can be estimated
using modeling, whereas field monitoring data may be required for the highest
tier. In the lower tiers, EECs are determined based on fate and transport
data (for aquatic systems) and residue chemistry and monitoring data (for
terrestrial systems), as well as the rate, frequency, timing, and method of
pesticide application. Exposure pathways are via water (for aquatic species)
and via food (vegetation and nontarget insects) for terrestrial species. EECs
in aquatic systems are estimated using hydrologic computer models developed by
EPA (EXAMS and SWRRB). These models estimate minimum and maximum
concentrations over time.
EECs for the first tier in terrestrial systems are derived using
pesticide residue profiles based upon published monitoring data. This method
uses the maximum and mean estimates of pesticide concentration immediately
following application. Different estimates of residue concentrations are
derived for different types of vegetation and nontarget insects. The residue
estimates are the EEC and are developed for those types of vegetation and
insects believed to comprise the diet of the nontarget species. EECs are
compared to RRCs derived from dietary concentration dose-response data. If
only dosage data are available, these values are converted into dietary
concentration values using assumed body weight and food intake values for the
nontarget species.
Risk Characterization
Risks are estimated by comparing the EEC to a set of regulatory risk
criteria (RRC). The RRCs for acute toxicity are equal to the LCjQ or U>50
divided by a safety factor of either 5, 10, or 20. The safety factors were
derived by examining a cross-section of existing acute lethality dose-response
data and determining the fraction of the median lethal value that corresponds
to mortality in 0.1% of the population. (Mortality of 0.1% was regarded as
sufficiently protective of a population.) For the typical (average)
dose-response curve, a value one-fifth of the median lethal value corresponds
to mortality in 0.1% of the population. Hence, a safety factor of five is
applied to the acute toxicity value to derive an RRC. An additional safety
factor of two is used for aquatic species because, it was reasoned, these
species are less capable of limiting their exposure to contaminants through
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SEP for Ecological Risk EPA/OPP 1986 Page C-44
TABLE C4-1
WILDLIFE AMD AQOATIC ORGANISM DATA REQUIREMENTS
Avian and Mammalian Tests
Avian single-dose oral LD50
Avian dietary LC5Q
Wild mammal toxicity
Avian reproduction
Simulated and actual field testing
Honey bee: acute contact U>5o
Honey bee: toxicity of foliar residues
Aquatic Organism Tests
Freshwater fish acute toxicity
Freshwater invertebrate acute toxicity
Estuarine/marine organism acute toxicity
Fish early life-stage study
Aquatic invertebrate life-cycle study
Aquatic organism bioconcentration
Simulated or actual field testing
Source: 40 CFR 157.145 and EPA/OPP 1986.
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SEP for Ecological Risk EPA/OPP 1986 Page C-45
food switching or by moving out of treated areas (e.g., ponds). Safety
factors of 10 and 20 are used for terrestrial and aquatic endangered species,
respectively. These higher safety factors are used for endangered species
because even a single death in these species is considered to be of special
concern. There is no safety factor applied to chronic no-effect-level
toxicity values to account for the uncertainty associated with laboratory-co-
field extrapolations.
If the EEC exceeds the RRC, a risk is presumed to exist and further
testing may be required. The degree of risk is not addressed. Using this
approach, potential risks to aquatic and terrestrial organisms following
exposures to a single chemical via a single route of exposure are evaluated;
multiple chemical exposures and multiple pathways are not considered, and,
therefore potential antagonistic, synergistic, or potentiating interactions
are not considered.
4.1.3 Operational Resource Requirements
The operational resource requirements of the Standard Evaluation
Procedures approach should be low to moderate for chemicals for which toxicity
information already exists. Operational resource requirements will be much
higher for new chemicals for which testing will be required. The approach
would be relatively simple to implement.
4.1.4 Stimnpary
The Standard Evaluation Procedures for evaluating ecological risks
proposed by EPA's Environmental Effects Branch is a quantitative risk
evaluation approach. The primary advantages of this approach are that ic is
relatively simple to use. In the first tier assessment, it uses acute
toxicity data that are readily available for many chemicals. In addition,-
this approach considers the characteristics of potential receptors when
selecting indicator species and determining terrestrial exposures. To
extrapolate to effects of chronic exposure from acute toxicity data, acute
LD5Q or LC50 values are divided by safety factors. A safety factor of 5 is
used for acute data from birds and mammals, while a factor of 10 is used for
aquatic organisms because they cannot avoid exposure as easily. An additional
safety factor of 2 Is applied if an endangered species might be at risk. Mo
safety factor is applied to chronic no-effect-levels determined in the
laboratory.
The Standard Evaluation Procedures, however, have several limitations.
Many of these limitations are inherent weaknesses of the quotient method;
others are unique to this method. Some of the major limitations of this
approach are:
The method does not compensate for differences between laboratory
and field populations.
It cannot be used for estimating indirect effects of toxicants, such
as food chain bioaccumulation.
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SEP for Ecological Risk EPA/OPP 1986 Page C-46
It has unknown reliability (i.e., has not been validated) and does
not quantify uncertainties.
It does not account for other ecosystem effects (e.g., predator-prey
relationships, community production/metabolism, structural shifts,
etc.).
It does not consider factors (biotic and abiotic) that might
influence toxicity.
Only one exposure pathway is considered for aquatic species and one
for terrestrial species.
Spatial variation in contaminant concentration is not considered.
Multiple chemical exposures-are not considered. While this is
sufficient for evaluating the potential effects of application of a
single pesticide, it would not be applicable for situations where
concomitant exposure to more than one chemical might occur.
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CMRA Onishi et al. 1982, 1985 Page C-47
4.2 COMPUTER-BASED ENVIRONMENTAL EXPOSURE AND RISK ASSESSMENT
METHODOLOGY FOR HAZARDOUS MATERIALS (CHEMICAL MIGRATION
RISK ASSESSMENT; Onishi et a!- 1982, 1985)
4.2.1 Introduction
The Chemical Migration Risk Assessment (CMRA.) methodology has two stated
objectives as described by Onishi et al. (1982): (1) to predict the occurrence
and duration of pesticide concentrations in surface waters receiving runoff
from agricultural lands, and (2) to develop a preliminary risk assessment
procedure to predict the potential damage to aquatic biota. Although these
were the stated purposes, the methodology can be used with little or no
modification to evaluate the impact of any organic chemical which enters water
from land.
The methodology is actually a series of models developed for EPA Athens
Laboratory by Batelle Northwest which may be used in conjunction with other
models developed by the U.S. Department of Agriculture or Nuclear Regulatory
Commission. The algorithms have also been -incorporated in the Center for
Exposure Assessment Modeling (CEAM) - supported model HSPF (Hydrologic
Simulation Program - FORTRAN). The basic architecture of the methodology i's
shown in Figure C4-1. Individual models from the methodology can be linked
together in various fashions depending on the type of water body and
dimensionality of the system. The methodology has been documented extensively
(Table C4-2); however, the most important overview documents are Onishi et al.
(1982 and 1985).
4.2.2 Description of Method
Receptor Characterization
The methodology is more hydrologically than biologically oriented and
thus is weak in its receptor characterization components. It can consider
freshwater, estuarine, or marine systems. Since the output is a probabilistic
risk assessment, it can deal with both individual or population risks if the
size of the population is known. The operator has considerable flexibility in
choosing receptors and time frames. For example, risk to an early life stage
(ELS) of a particular species could be assessed if toxicity data and
sufficient information concerning the space-time location of the ELS were
known.
Hazard Assessment
The model basically integrates acute (LC50) and chronic (MATC) values
into a time-dose-response curve. Only three endpoints, "safe", "mortality",
and "sublethality" are used, although these could be expanded. No toxicity
modifying factors are considered. Uncertainty is considered implicitly by
using safety factors in developing MATCs or extrapolating between 1X50 values
obtained under different time frames. Many of the apparent problems with the
hazard assessment component (e.g., failure to consider other endpoints or che
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CUBA
Onishi ec al. 1982, 1985
FIGURE C4-1
SCHEMATIC DIAGRAM OF THE CMRA METHODOLOGY
(Adapted from: Onishi e£ a4. 1985)
Page C-48
INPUT DATA
Terrestrial:
ANALYSIS
Meteorological information
Contaminant data
application rate
physical properties
Watershed characteristics
Overland Contaminant
Transport Modeling
Aquatic:
Channel characteristics
Sediment characteristics
Containment properties
Boundary conditions
flow
sediment
containment
Surface Water
Contaminant Modeling
Contaminant Concentrations
Aquatic Toxicological Data:
LC50
MATC
1
Risk Assessment
T
Estimates of Lethal and
Sublethal Damages on
Aquatic Biota
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CMRA Onishi et al. 1982, 1985 Page C-49
TABLE C4-2
SELECTED BIBLIOGRAPHY
Onishi, Y., Olsen, A.R., Parkhurst, M.A., and Whelan, G. (1985) Computer Based
Environmental Exposure and Risk Assessment Methodology for Hazardous
Materials. Jour. Hazardous Materials 10: 389-417.
Onishi, Y., Brown, S.M., Olsen, A.R., Parkhurst, M.A., Wise, S.E., and
Walters, W.H. (1982) Methodology for Overland and Instream Migration and Risk
Assessment of Pesticides. EPA-600/3-82-024.
Onishi,Y. and Wise, S.E. (1982) User's Manual for the Instream Sediment -
Contaminant Transport Model Seratra. EPA-600/3-82-055.
Onishi, Y. and Wise, S.E. (1982) Mathematical Model, SERATRA. for Sediment
Contaminant Transport in Rivers and its Application to Pesticide Transport in
Four Mile and Wolf Creeks in Iowa. EPA-600/3-82-045.
Olsen, A.R. and Wise, S.E. (1982) Frequency Analysis of Pesticide
Concentrations for Risk Assessment (FRANCO Model).
EPA-600/3-82-044.
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CMRA Onishi et al. 1982, 1985 Page C-50
lack of inclusion of a true exposure-response curve) are not inherent in the
methodology--given more data and users who are biologically oriented, several
of the apparent problems could be overcome.
Exposure Assessment
The methodology excels at exposure assessment since many of the component -
models were developed specifically for this purpose. The user can select from
a wide range of hydrologic solute transport models both in overland flow and
in surface water. If the hydrodynamics are complex, it is relatively simple
to link a hydrodynamic model with the rest of the methodology. The output is
typically stochastic although it could be forced to be deterministic. In a
typical application, this output is in the form of the probability of
occurrence and duration of specific toxicant concentrations; thus there is a
de facto consideration of uncertainty. The exposure component has been
verified with two pesticides: toxaph'ene and Alachlor. Various models used in
the exposure component have been validated by comparison of field results to
predicted values; however, the entire component has not been validated. If
required, the exposure models can be calibrated by manipulation of input
variables.
Risk Characterization
CMRA uses probabilistic risk assessment to evaluate impacts on aquatic
life. In this component, the model known as FRANCO takes the time-dependent
output of the exposure models and calculates the number of times and fraction
of the time that concentration-duration levels (e.g., 96 hour LC50) are
exceeded. Risk is then calculated as the frequency of occurrence of an event
(e.g., exceeding an MATC) and its consequences (e.g., mortality). Certain
assumptions are used in applying this method, including: a 96 hour cutoff
between acute and chronic; concentrations less than the MATC are "safe";
concentrations between the MATC and the concentration-duration curve result in
"sublethality" or "damage"; and concentrations over the concentration-duration
curve result in "mortality". If the size of the population is known, this
leads to prediction of the number of effects. The probabilities shown on the
output can also be interpreted as risk to an individual. If toxicological
data on a chemical mixture are available, then the effect of the mixture can
be assessed. As it stands, the methodology is limited to exposure through
water. Food chain exposure is not considered.
4.2.3 Operational Resource Requirements
These depend on which of the overland transport and/or surface water
solute transport models are selected. A mini/mainframe would be required to
have the flexibility to use all model components. Since the methodology is
most sophisticated with respect to exposure modeling, an experienced
hydrologist probably would be required to run the models and interpret the
results.
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CMRA Onishi et al. 1982, 1985 Page C-51
4.2.4 <;"BP'ftry
Both the overland transport and surface water solute transport components
were developed independently of the risk assessment (FRANCO) component. The
aquatic toxicology input is also independent. Therefore, use of the
methodology should not be considered to be restricted to the models originally
available (e.g., TODAM, SERATRA, FETRA, FLESCOT or Mixed Tank Models). The
strongest component of the methodology is the FRANCO model, a probabilistic
risk assessment which has considerably more utility and power than
methodologies using quotients. It appears to be quite flexible and adaptable
to terrestrial receptors, mixtures of toxicants, and even ecosystem effects
given sufficient information. The overall methodology has been verified for
two pesticides: coxaphene and Alachlor; it has not been validated, however.
Although it presently considers as endpoints only "safe", "mortality", and
"sublethaiity", it could be expanded-to include such parameters as impacts on
target organs or specific disease states if the toxicological data were
available.
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Sensitive Environments EPA/OSW 1987a Page C-52
5.0 METHODS/ASSESSMENTS DEVELOPED UNDER RCRA
5.1 POTENTIAL FOR ENVIRONMENTAL DAMAGE: PROXIMITY OF NINE SITES TO
SENSITIVE ENVIRONMENTS (EPA/OSW 1987a)
5.1.1 Introduction
ICF Incorporated has developed for EPA's Office of Solid Waste a
qualitative approach for ecological assessment that is based on the proximity
of sensitive environments to potential sources of contaminants. The approach
is most appropriate for screening-level analysis, and has been used to
preliminarily assess the potential for environmental impacts from oil and gas
and mining activities.
The term sensitive environment refers to environmental areas that are
either ecologically critical, unique,-or vulnerable; are of particular
cultural significance; or are set aside for the purpose of conservation. The
assessment approach is based on the concept that actual or potential impacts
to sensitive environments are generally of greater concern to society than
comparable impacts in other environmental areas. Waste sources near, sensitive
environments are regarded as having a greater potential for environmental
impact. Although the proximity of sensitive environments to. waste sources--Is
not an explicit criterion for determining ecological risk, it is an important
consideration.
The approach is not a complete ecological risk assessment methodology
because exposure and hazard assessments are not conducted. It does, however,
provide potentially useful concepts for receptor characterization and risk
determination.
5.1.2 Description of Method
Receptor Characterization
Four categories of sensitive environments were identified, encompassing
both aquatic and terrestrial areas: endangered and threatened species
habitats; wetlands; National forest system lands; and National park system
lands. For any particular geographic area, sensitive environments can be
identified using maps or available data bases. For example, information
published by the U.S. Fish and Wildlife Service (USFWS) can be used to
identify endangered and threatened species habitats. The USFWS has identified
the historical ranges of approximately 400 endangered and threatened species
and the critical habitats for 96 species. Also, the Nature Conservancy
maintains a data base that can be used to identify endangered and threatened
species habitats for any given location. Wetlands can be identified using
U.S. Geological Survey 7.5 minute quadrangle maps or quadrangle maps published
by the National Wetlands Inventory (USFWS). National forests and parks also
can be located on quadrangle maps.
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Sensitive Environments EPA/OSW 1987a Page C-53
Risk Characterization
Under this assessment scheme, the degree of risk posed by any contaminant
source is qualitatively assessed using the spatial relationship between the
source and the receptor (sensitive environment). The potential for
environmental damage increases as the distance to the sensitive environment
decreases or as the number of proximate sensitive environments increases.
5.1.3 Operational Resource Requirements
This approach is based on information that is readily available for
virtually all areas of the United States. The effort and costs for obtaining
and interpreting the information are low.
5.1.4 Summary
This approach is an easily-implemented, low-cost, qualitative method for
ranking the potential for waste sites to impose adverse impacts on valuable
communities and portions of ecosystems. The potential for. impacts is
determined qualitatively based on the proximity of sensitive environments tp a
contaminant source. Because no absolute criteria relating distance to risk
can be established without consideration of waste characteristics and exposure
pathways, the approach is limited to use as a comparative screening analysis
where several contaminant sources are being evaluated. The approach is not
precise and has numerous other limitations (e.g., no hazard or exposure
assessments) that make it applicable only to situations requiring rough
screening analyses.
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HWT Risk-based Variance EPA/OSW 1987 Page C-54
5.2 VARIANCE FROM THE SECONDARY CONTAINMENT REQUIREMENTS OF HAZARDOUS
WASTE TANK SYSTEMS: VOLUME II: RISK-BASED VARIANCE (EPA/OSW 1987b)
5.2.1 Introduction
IGF Inc. prepared guidelines for EPA's Office of Solid Waste to estimate
environmental risks posed by the release of waste constituents from hazardous
waste tanks. The approach is based on the work of Barnthouse et al. (reviewed
in Section 7.1.1 of this appendix) and EPA/OPP (reviewed in Section 4.1 of
this appendix). The approach uses the quotient method in which estimated
environmental concentrations (EEC) are compared to reference concentrations
(RC) believed to be protective of aquatic and terrestrial species. Estimates
of environmental concentrations can be obtained using computerized fate and
transport models or simple mass-balance equations. The toxicity-based RCs are
EPA water quality criteria, or LOELs or NOELs divided by safety factors.
The ratio of the EEC to the RC for all chemicals are summed to produce a
hazard index, and the index is compared to predetermined concern levels. A
hazard index less than or equal to 1.0 indicates a low probability of .
environmental harm. A value between 1.0 and 10 is assumed to be indicative of
possible effects. A value greater than or equal to 10 indicates probable
environmental harm. If the hazard index is less than 1.0, the environmental
impact evaluation is considered complete. If the hazard index is greater than
10, an environmental site evaluation must be conducted.
5.2.2 Description of Method
Receptor Characterization
Both aquatic and terrestrial receptors are considered under this method.
Assessments are focussed at the population level, although the value of an
individual member of an endangered species is considered by using larger
safety factors for toxicity values for terrestrial endangered species. The
species potentially occurring at a site are identified and differentiated
based on potential differential sensitivity to toxicants. Toxicity values
based on indicator species which most closely resemble the sensitivity of
potential receptors are selected for the risk assessment. Other
characteristics of potential receptors are not considered.
Hazard Assessment
Hazard assessments are conducted at the population level, and acute and
chronic toxicity thresholds are used to calculate RCs; dose-response data or
community and ecosystem level responses are not considered. For aquatic
toxicity, the reference concentrations are set equal to EPA's water quality
criteria (WQC). If a WQC is not available for a given chemical, a chronic
LOEL divided by 10, or an acute LOEL divided by 100 (whichever is lowest), is
used for hazard assessment. The factors of 10 and 100 were derived from
statistical and other analyses of available toxicity data and are intended co
incorporate uncertainties due to test species sensitivity, acute to chronic
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HWT Risk-based Variance EPA/OSW 1987 • Page C-55
toxicity extrapolations, and field to laboratory differences. If the WQC for
a given chemical is not relevant for species or conditions at the site,
site-specific criteria can be derived by recalculating a WQC or by conducting
additional tests using water and/or species from the site. For terrestrial
species, the lowest of the identified acute NOELs divided by 100 or the lowest
of the identified chronic NOEL divided by 10 is used in the hazard
evaluation. An additional safety factor of 10 is used when evaluating
potential impacts to endangered species. The rationale for the terrestrial
safety factors was not provided.
Exposure Assessment
Specific procedures for estimating environmental exposure concentrations
are not provided, although procedures ranging from simple mass balance
equations to complex models are proposed. Exposures via water (for aquatic
species), and via food (for terrestrial species) are the exposure pathways
considered. Procedures used by EPA's Office of Pesticide Programs (EPA/OPP
1986, Section 4.1) are used to derive exposure doses from exposure
concentrations in terrestrial systems. Abiotic conditions (e.g., pH,
alkalinity, suspended solids, salinity) which may modify aquatic toxicity are
considered during the exposure assessments. If site conditions are /•
sufficiently different from the test conditions under which the toxicity
criteria were derived, site-specific criteria may be developed.
Risk Characterization
Risks are estimated by comparing the EEC to the RC to obtain a ratio.
The ratios are then summed to provide a hazard index. If the hazard index is
less than or equal to 1.0, a low probability of environmental harm is assumed.
A value between 1.0 and 10 is indicative of possible harmful effects. A value
greater than or equal to 10 is an indication of probable harmful effects.
This approach assumes strict additivity of doses (i.e., similar modes of
toxicity, although chemical potency can vary). Thus, antagonistic,
synergistic, or potentiating interactions are not considered.
5.2.3 Operational Resource Requirements
The level of effort required for this approach should be low to moderate
for chemicals for which toxicity information already exists. Operational
resource requirements will be much higher if site-specific criteria are
required.
IFor a given species and chemical, a NOEL is identified as the highest
exposure dose or concentration tested that does not produce an observable
response. Identification of a NOEL requires identification of a LOEL (lowest-
observed-effect-level) at a higher dose or concentration.
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HWT Risk-based Variance EPA/OSW 1987 Page C-56
5.2.4
The Risk-Based Variance Procedure for evaluating ecological risks posed
by Che release of waste constituents from hazardous waste tanks is a semi-
quantitative risk evaluation approach. The approach is similar to the
quotient method proposed by Barnthouse et al. (1986; Section 7.1.1). The
primary advantages of this approach are that it is relatively simple to
implement and that it uses data that are readily available for many chemicals.
In addition, this approach considers site characteristics, including receptor
characteristics, when selecting or deriving appropriate reference
concentrations. This approach has several limitations, however. Many of
these limitations are inherent weaknesses of the quotient method; others are
unique to this method:
The method does not account"for effects of incremental dosages
(i.e., exposure - response).
The method does not compensate for differences between laboratory
and field populations.
It cannot be used for estimating indirect effects of toxicants, such
as food chain interactions.
It has not been validated and does not quantify uncertainties.
It does not account for other ecosystem effects (e.g., predator-prey
relationships, community production/metabolism, structural shifts,
etc.).
Only one exposure pathway is considered for aquatic species and one
for terrestrial species.
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RCRA Risk-Cost Model EPA/OSW 1984 Page C-57
5.3 THE RCRA RISK-COST ANALYSIS MODEL (EPA/OSW 1984)
5.3.1 Introduction
Subtitle C of RCRA authorizes EPA to develop a national regulatory
program for the management of hazardous wastes. As part of the regulatory
program, EPA must provide standards for hazardous waste treatment, storage and
disposal facilities. The purpose of the RCRA risk-cost analysis model,
developed by IGF Inc., is to provide a consistent method to analyze the
relative risks and costs of different waste management practices.
Conceptually, the model can be regarded as a three-dimensional matrix,
each cell of which is a combination of a waste stream, an environment, and a
waste management practice or technology. The model then calculates the risks
and costs associated with each waste/environment/technology combination.
Absolute risk values are not calculated in the model, only a relative scale is
used. Although the model includes modules to assess risks to human health as
well as risks to ecosystems, this summary only addresses the portion of the
model that assesses ecological risks.
5.3.2 Description of Method
Receptor Characterization
Risks to both terrestrial and aquatic (surface water) receptors are
assessed. The United States is divided into 559 areas based on three-digit
zip codes. Each zip code area is defined as having one of seven possible
surface water environments (large, medium, or small river, rivers with large
drainage but low flow, marshes, seacoasts, and lakes). In addition, the zip
code area is designated as being "more important" if it contains a National
park, seashore, wilderness area, monument, river, lake shore or historic site,
commercial fishing area, or public or private beach.
Hazard Assessment
Although the model addresses the risks associated with waste streams, the
hazard assessment is conducted only for the most toxic constituent of the
waste stream. Toxicity is assessed by constructing an ecosystem dose-response
curve for the most toxic constituent. This curve is based on four case
studies where minimal to catastrophic effects were documented. The curve
derived from these case studies is shown in Figure C5-1 and is based on two
assumptions (1) the ecosystem threshold is equal to the threshold
concentration for the most sensitive species; (2) the range of concentrations
between the threshold and catastrophic level is two to three orders of
magnitude (depending on how "catastrophic" is defined, and how carefully the
threshold is measured). The model's ecosystem damage function varies between
0 (no damage) when the concentration of the pollutant is at or below a
threshold and 1.0 (i.e., 100% damage) at pollutant concentrations 2.5 orders
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RCRA Risk-Cose Model
EPA/OSW 1984 page c_58
FIGURE C5-1
THE ECOSYSTEM DAMAGE FUNCTION
(Source: EPA/OSW 1984)
EXPOSURE-RESPONSE RELATIONS FOR SPECIES
AND DAMAGE FUNCTION FOR ECOSYSTEM a/
e ' e * e' e • e
CT V V V V
» Ambient chemical concentration in water
» Exposure-response relation for the i species.
* Threshold concentration for the i species.
.ch
D(C)
• Concentration giving a 50 percent response rate for the i
species.
* Concentration giving a 100 percent response rate for the i
species.
* Damage function for the ecosystem.
* Threshold concentration for the ecosystem.
* Catastrophic or fatal concentration for the ecosystem
.th
a/ Assumes a given chemical.
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RCRA Risk-Cost Model EPA/OSW 1984 Page C-59
of magnitude higher. The curve is linear with log concentrations. The lowest
effect level for an individual organism is defined and used as a surrogate to
estimate the extent of ecosystem level damages.
The ecosystem threshold value is derived separately for salt water and
freshwater aquatic species; subcategories within these receptors are not
addressed. Hazard to terrestrial systems is assessed by analogy to aquatic
systems because of the lack of appropriate toxicity and exposure data for
terrestrial ecosystems. For most of the chemicals, the threshold ecosystem
value is chosen as the lowest of either salt water or freshwater species
threshold concentrations. A different procedure is used for chemicals that
bioconcentrate extensively (e.g., chlordane, hexachlorobenzene, PCB-1254,
toxaphene, and mercury). For these chemicals, the dose-response curve is
adjusted downward; the catastrophic damage concentration is set by analogy to
the aquatic systems, but the ecosystem threshold is assumed to be 5 orders of
magnitude below the catastrophic concentration on the basis of two case
studies.
Uncertainty in deriving the aquatic threshold values is addressed by
applying a series of safety factors to the lowest experimental threshold •
values. The safety factors are based on data quality, data completeness, .
exposure duration, and the absence of experimental threshold. The safety1
factors for data completeness vary from 1 to 10 and depend on the quantity and
representativeness of the available data for a given chemical. If only acute
data are available, empirically derived "acute-chronic ratios" for
structurally similar chemicals are used. If the lowest concentration at which
a chemical was tested still produced toxic effects, a threshold concentration
is estimated by dividing the lowest effect concentration by 10. Default
values, based on structurally similar chemicals, are used for chemicals for
which no data are available.
Exposure Assessment
For the aquatic exposure assessment, surface water concentrations are
estimated with distance downgradient from the point of release. A simple
steady-state model is used, based on the release rate of the chemical, the
flow rate of the water body, and overall decay rates. Release rates of
chemicals are based on the treatment technology specified and include both
continuous and intermittent releases. Exposure to terrestrial systems is
assumed to be equivalent to the aquatic component of the environment, although
insufficient justification is given for this approach.
Risk Characterization
A score is calculated for each cell of the waste
stream/environment/treatment technology matrix. The first stage score is
calculated by integrating the aquatic and terrestrial ecosystem damage
functions over distance downstream from the release. The first stage score is
the log^Q of the integrated damage. The first stage scores then are modified
upward by an order of magnitude if an "important environment" exists within
the zip-code area. Aquatic and terrestrial scores are added to obtain the
final score.
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RCRA Risk-Cost Model EPA/OSW 1984 Page C-60
Although the score is relative, parts of the model, such as the ecosystem
damage function or the exposure model, could be calibrated and validated. As
knowledge of ecological effects increase, other damage functions or exposure
models could easily be incorporated into this framework.
5.3.3 Operational Resource Requirements
The chemical-specific data requirements for this model are low and are
readily available for many chemicals. The model development effort is
complete, and the model can be run with a relatively low level of skill. As
the model exists now, the ecosystem component cannot be run separately; a
mainframe computer is needed to run the complete model.
5.3.4 Summary
The ecological component of the RCRA Risk-Cost Analysis Model scores the
relative risk associated with different treatment technologies, waste streams,
and environments. One strength of the model is that it is based on empirical
studies of ecosystem responses, yet operates using readily-available data. It
incorporates a measure of the areal extent of predicted damage (e.g., stream
miles), and considers the presence of socially or biologically important
environments. It is very flexible and could easily be modified to incorporate
findings from recent or future studies. A limitation of the model is that
terrestrial risk is assessed almost totally by analogy to the aquatic system.
Because only the most toxic component of a waste stream is considered, it
would be difficult to use the model to obtain an absolute estimate of risk.
However, components of the model, such as the aquatic hazard assessment, could
be combined with other exposure assessments to estimate absolute risk for
individual chemicals.
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Estimating Concern Levels EPA/OTS 1984 Page C-61
6.0 METHODS/ASSESSMENTS DEVELOPED UNDER TSCA
6.1 ESTIMATING -CONCERN LEVELS" FOR CONCENTRATIONS OF CHEMICAL SUBSTANCES
IN THE ENVIRONMENT (EPA/OTS 1984)
6.1.1 Introduction
The purpose of this method is to provide a framework for calculating
acceptable release levels of chemicals under the Premanufacturing Notification
(PMN) program of the Office of Toxic Substances. The method can be used to
identify concentrations of chemicals that might cause adverse environmental
effects in aquatic populations, and to identify chemicals which should be
tested more rigorously under Section 5 of TSCA.
A series of assessment factors are presented to apply to the lowest
observed effect level of a given chemical. Guidelines are presented to help
the assessor choose the appropriate assessment factor, based on the amount and
type of toxicological data available.
6.1.2 Description of Method
Receptor Characterization
The method has been developed using aquatic toxicity data and is intended
to predict chronic effects at the population level. Community effects are
assumed to mirror individual organism effects. The proposed approach is
generic and no specific receptors are identified. Although temporal
characteristics of receptors are not included, the assessment factors were
developed assuming that early life stages are more sensitive than adult life
stages. It also was assumed that natural conditions are less conducive to
survival than laboratory conditions.
Hazard Assessment
The concern level is derived by applying an assessment factor to the
lowest observed effect level for a selected chemical. The assessment factor
is chosen to reflect the degree of extrapolation from the available toxicity
data to data needed to assess effects in natural populations; assessment
factors of 0.1, 0.01, or 0.001 are multiplied by (1) the lowest chronic effect
concentration, (2) the lowest LC5Q concentration of many acute tests, or (3)
one LC5Q from an acute test or quantitative structure-activity relationship,
respectively. If there is in situ evidence of population effects from
biological monitoring or full scale field studies, the assessment factor
equals one. OTS's method of estimating levels of concern is intended to
identify concentrations of chemicals that, if met or exceeded, might cause
adverse environmental effects in populations at least 95% of the time. The
possibility that adverse effects might occur below the concern level is not
addressed.
The assessment factors were developed on the basis of statistical
analysis of laboratory toxicity tests, considering the relationships between
acute and chronic toxicity, laboratory test conditions and field conditions.
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Estimating Concern Levels EPA/OTS 1984 Page C-62
and different species sensitivities. Each assessment factor was selected to
approximate a 50% confidence limit. In other words, applying the factor to a
given toxicity value (e.g., one acute LC5Q test) should encompass the toxicity
value of interest (e.g., LC5Q for the most sensitive species) 50% of the cime.
Toxicity of chemicals that have not been tested can be extrapolated from
appropriate analogs using Quantitative Structure Activity Relationships
(QSARs).
Exposure Assessment
An exposure assessment is not included as part of this method.
Risk Characterization
The methodology is designed to determine criteria for triggering concern.
It is a quantitative estimation of maximum contaminant levels for individual
chemicals in water that, if met or exceeded, might adversely effect aquatic
organisms. The method does not consider any toxicity modifying factors or
interactions with other contaminants. Statistical uncertainty is incorporated
into the assessment factors. The method could be field validated.
:
6.1.3 Operational Resource Requirements
The method is simple and can be used with a wide variety of chemicals.
The operational costs and effort are inversely proportional to amount of data
available for a given chemical. If little or no input toxicity data are
available, a high level of expertise on chemical property estimation methods
is necessary.
6.1.4 Summary
Under this approach, an assessment factor is applied to available
toxicological data to determine an environmental contaminant concentration
that is likely to result in adverse effects to aquatic organisms. The
strengths of the method are that it is very simple and that a hazard
assessment can be performed even with very limited toxicity data. Limitations
of the method include its failure to consider chemical and biological
interactions, or to use exposure-response data..
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Ecorisk in OTS EPA/OTS 1987 Page C-63
6.2 ECOLOGICAL RISK ASSESSMENT IN THE OFFICE OF TOXIC SUBSTANCES, PROBLEMS
AND PROGRESS (EPA/OTS 1987)
6.2.1 Introduction
The Office of Toxic Substance's Environmental Effects Branch (EEB) has
developed an approach for assessing the potential ecological hazards and risks
associated with new chemicals. This approach continues the work of Barnthouse
e_£ al. (1982, 1986) (reviewed in Section 7.1 of this appendix) and also
previous work in OTS (reviewed in Section 6.1 of this appendix). The purpose
of the ecological evaluation is to screen chemicals to identify those for
which additional testing is needed based on the potential for ecological risk.
The approach is a modified quotient method in which estimated environmental
concentrations associated with production, use, and disposal of a chemical are
compared to toxicity criteria modified by uncertainty assessment factors.
Ratios which exceed one indicate a potential for ecological impact and thereby
substantiate the need for further testing under the New Chemical Review
Process.
6.2.2 Description of Method
•.
Receptor Characterization
Theoretically, both aquatic and terrestrial receptors can be considered
under this method, but potential impacts to aquatic systems are emphasized
because the manufacture, use, and disposal of toxic chemicals most often
results in discharges to aquatic environments. Assessments are focused at the
population (species) level, and receptor populations are selected based on
their importance as commercial, sport, recreational, aesthetic or other
resources valued by society. Surrogate species are used in the quantitative
evaluations of risk. Surrogate species are selected based on the general
laboratory testing protocol developed by EEB and include fish, daphnids, and
algae. Seasonal or life-cycle characteristics and niche characteristics of
the receptors are not considered in the evaluation.
Hazard Assessment
Hazard assessments are conducted at the population level, and changes in
growth, development, mortality and reproduction are identified as the
endpoints of concern. Two phases of hazard assessments are conducted. The
first phase is a conceptual approach in which a fault tree analysis is used to
identify ecologically important impacts on populations (e.g., Figure C6-1).
Potential effects on growth, development, reproduction, and mortality are
classified as either direct effects or indirect effects (e.g., changes in
predator prey relationships, competition, habitat). These are further divided
into effects due to natural causes (e.g., climate, normal population
variations) and those due to toxic chemicals. Under the latter category,
acute and chronic effects are considered. Toxicity modifying factors, both
biotic (e.g., competition) and abiotic (e.g., climate change) are considered.
This approach provides a conceptual framework only, and therefore does not
specify the type of toxicity data to be used in hazard assessment or the
method by which to derive a hazard value.
-------
Ecorisk in OTS
EPA/OTS 1987
FIGURE C6-1
Page C-64
(Source:
EXAMPLE OF A FAULT TREE ANALYSIS
Barnthouse eg al. 1986 as used in EPA/OTS 1987)
/ N. MttMVMV^
40MNMTAT.O* / """V
-------
Ecorisk in OTS EPA/OTS 1987 Page C-65
The second phase of the hazard assessment involves the derivation of a
"concern level", which is a chemical concentration that if equaled or exceeded
would justify further testing of a chemical. Toxicological endpoints are
defined by the results of acute, chronic, or subchronic tests and are reported
as LC5QS, EC5QS, and MATCS. Toxicity values for the most sensitive species
are used to determine concern levels. Concern levels are derived by dividing
the toxicity value (LC5Q, ££50- or MATC) for a new chemical or an analog by an
"assessment factor." Assessment factors range from 1 to 1000 by multiples of
10, and are based on the degree of uncertainty associated with the toxicity
values (reviewed in Section 6.1 of this appendix). An assessment factor of 1
is used when the toxicity value is based on the results of a field test, and
an assessment factor of 1000 is used when there are only two or three
laboratory tests available for the new chemical or an analog, or if a
toxicological value is derived from a Quantitative Structure Activity
Relationship.
Exposure Assessment
Specific procedures for exposure assessment are not specified in the
method, but contaminant concentrations in water resulting from the production,
use, or disposal of a particular TSCA chemical must be estimated or measured.
Risk Characterization
Environmental concentrations that result or could result from simulated
or-actual conditions of chemical production, use, and disposal are compared to
the "concern level." Ratios that exceed 1 indicate the potential for
ecological impact and justify the need for further testing.
Using this approach, the potential for adverse effects on aquatic
organisms following exposures to a single chemical via a single route of
exposure are evaluated; multiple chemicals and multiple pathways are not
considered. The method is broadly applicable as a screening methodology to
identify areas for further study. The method, as currently applied, does not
address uncertainty in the exposure or risk estimate. Uncertainty in the
hazard assessment is addressed via the use of assessment factors. The method
has not been validated, field-tested, or calibrated.
6.2.3 Operational Resource Requirements
The resource requirements of the EEB risk assessment method are low. The
approach is relatively simple to implement and requires minimal data, most of
which is usually readily available for the chemical being evaluated or an
analog.
6.2.4 Summary
The method currently used by EEB for identifying chemicals for further
study is a modified quotient method. The primary advantages of this approach
are that it is easy to implement and has minimal data requirements.
Consequently, it is an approach useful for the rapid screening of new
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Ecorisk in OTS EPA/OTS 1987 Page C-66
chemicals. There are, however, several limitations associated with this
approach, and most of these limitations are inherent limitations of the
quotient method: it does not routinely take into account exposure-response;
it does not consider indirect effects; it has no predictive capabilities; and
it addresses the uncertainty associated with variation in taxonomic and life-
stage sensitivities with generic application factors rather than factors
derived specifically for a chemical group or for species in specific regions.
EEB has recognized these deficiencies and has suggested additional
approaches for consideration in ecological assessments. The goal of EEB is to
be able to assess both direct and indirect effects, and eventually to evaluate
potential for recovery. The Ecosystem Uncertainty Analysis (EUA) of
Barnthouse e_£ al. (1982, 1986; see Section 7.1) is regarded by EEB as a
potential tool for evaluating the potential impacts of indirect toxic effects
(e.g., on other trophic levels) as well as direct toxic actions on
populations. The model used to illustrate the EUA methodology, SWACOM,
attempts to predict the effects of a toxicant on a pond ecosystem. Included
in the assessment are estimates of direct mortality and indirect changes in
population, community, and ecosystem structure. The Risk Analysis and
Management Alternatives (RAMAS) approach is regarded as a potentially viable
method of addressing the effects of a toxicant on the age/size class of any.
population. RAMAS estimates direct toxic action on the life stages of various
populations by the use of a Monte Carlo simulation of age structured
populations. It is designed to address the probability that a population will
fall below a given threshold within a specified time, and also predicts the
number of individuals within a given age class. EEB is currently evaluating
the potential utility of these and other methods.
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Users Manual for Eco. Risk Bartithouse et al. 1982, 1986 Page C-67
7.0 OTHER METHODS/ASSESSMENTS
7.1 USERS MANUAL FOR ECOLOGICAL RISK ASSESSMENT (Barnthouse et al. 1982,
1986)
7.1.1 THE QUOTIENT METHOD
7.1.1.1 Introduction
The quotient method (QM) (Barnthouse et al. 1986) compares an estimated
environmental concentration (EEC) to a toxicological benchmark (e.g., LC50,
GMATC, EPA AWQC). The purpose is either to (1) predict if an EEC is of
concern, or (2) rank a series of contaminants or contaminant sources by their
potential for producing adverse environmental effects. The procedures were
developed to support EPA's synfuels research program.
7.1.1.2 Description of Method
Receptor Characterization
The type of receptor selected under this approach depends on the type of
benchmark value used. The same type of toxicological benchmark (e.g., MATC'or
LC50) is required for all chemicals to make comparisons between quotients.
Moreover, a single representative benchmark value is required for each
chemical. This value could be the toxicological benchmark value for either an
indicator or sensitive species. Alternatively, a composite benchmark such as
an Ambient Water Quality Criterion, designed to be protective of the majority
of aquatic species, could be utilized.
In general, toxicological benchmarks (TBs) for aquatic indicator species
(population level) are used, but TBs for terrestrial species are available for
some types of exposure routes. The approach is usually generic, but a small
degree of specificity can be introduced by the selection of appropriate
indicator species. If exposures are expected to be brief, on the order of
days, either adult or sensitive life-stage benchmarks could be used as the
season dictates. For chronic exposures, LCls or MATC benchmarks are most
appropriate.
Hazard Assessment
There are two statistical types of toxicological benchmarks: (1) those
that prescribe an effect level (e.g. LC50 or LCI) based on a dose-response
relationship, and (2) those that are based on hypothesis testing. The second
type of benchmark is exemplified by the MATC which is assumed to lie between a
no-observed-effect level (NOEL) and a lowest-observed-effect level (LOEL). To
identify NOELs and LOELs, responses at exposure concentrations are compared
with control responses (no exposure) to test the null hypothesis that they are
the same as the control responses. Using conventional hypothesis testing
procedures that set alpha - 0.05 and leave beta unconstrained, one avoids
declaring that a concentration is toxic when it is not with a high degree of
certainty (type I error). However, there is an undefined chance of declaring
that a concentration is not toxic when it is (type II error). Thus.
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Users Manual for Eco. Risk Barnthouse e_£ aj.. 1982, 1986 Page C-68
coxicological benchmarks based on dose-response relationships are preferable
to those obtained by hypothesis testing.
If the quotient method is used primarily as a screening tool to rank
chemicals according to their potential for adverse environmental effects, the
quotients are compared directly. If the QM is used to provide an estimate of
environmental risk, uncertainty about the relationship of the TB to the
organisms and exposure situation at hand is treated by the application of
safety factors. Recently, application of a string of safety factors, each
representing a different source of uncertainty, has become common. The
underlying belief appears to be that "everything will go wrong at once", and
the result can be an extremely conservative evaluation.
Exposure Assessment
This method does not include arr exposure assessment.
Risk Characterization
"Risk" is based on the ratio of the EEC to the appropriate toxicological
criterion. For single chemical assessments, concern levels have been set
arbitrarily as follows: no concern, quotient < 0.1; possible concern, 0.1 <
quotient < 10; probable concern, quotient > 10. For multiple contaminant
screening, the quotients are ranked relative to one another.
7.1.1.3 Operational Resource Requirements
Computationally, the quotient method is extremely simple. However, the
data requirements can be extensive depending on the benchmarks and chemicals
under consideration.
7.1.1.4 Summary
The quotient method (QM) can help identify EECs that are unlikely to be
of concern. When the EEC exceeds the lowest toxicological benchmark
concentration, the QM cannot predict the degree or type of potential
environmental impacts. The method does not lend itself to consideration of
multiple routes of exposure, food chains, or any estimates of uncertainty.
The QM is useful mostly as an initial screening tool. With appropriate
professional judgment, the QM could provide a useful method of identifying
contaminants that are below levels of possible concern. If a few contaminants
must be selected to represent the waste stream of a particular source category
in more extensive ecological risk modeling, the QM can provide a useful first
tier screen. Even this limited application, however, can produce misleading
results if the quality of information available for the contaminants is
variable, or if the slope of the dose-response curve differs drastically
between chemicals.
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Users Manual for Eco. Risk Barnthouse e_£ al. 1982, 1986 Page C-69
7.1.2 ANALYSIS OF EXTRAPOLATION ERROR
7.1.2.1 Introduction
The analysis of extrapolation error (AEE) is a method for estimating the
probability that an environmental contaminant concentration (EEC) exceeds
measured or unmeasured toxicological benchmark values (TBVs). Unmeasured TBVs
are extrapolated from .available information. The uncertainty associated with
the extrapolation is quantified through regression analysis rather than by the
application of safety factors. The AEE method emphasizes that uncertainty in
estimations contributes to risk.
7.1.2.2 Description of Method
Receptor Characterization
The AEE provides a methodology for estimating unmeasured toxicological
benchmarks. As a consequence, receptor characterization is far more flexible
than with the previously discussed quotient method. The most extensive
toxicological data available to support statistical regression analyses are
for aquatic species. For short-term exposure scenarios, toxicological .
endpoints for adults or sensitive life stages can be chosen according to the
seasonality of the anticipated exposure. For chronic exposures, maximum
allowable toxicant concentrations (MATCs) or LCls are more appropriate.
Hazard Assessment
In the absence of appropriate laboratory test data for a species and
toxicological benchmark of interest, the benchmarks are extrapolated from
available toxicity data using regression analysis. The AEE hazard assessment
consists of four steps: (1) definition of the endpoint for the risk
assessment in terms of a toxicological endpoint (e.g., the probability of
exceeding the brook trout MATC); (2) identification of the existing datum most
closely related to the endpoint (e.g. a rainbow trout 96 hr LC50); (3)
identification of the necessary extrapolations (e.g. rainbow trout to brook
trout, 96 hr LC50 to MATC); and (4) the statistical regression analysis.
In step four, an errors-in-variables regression model provides the best
estimate of the unmeasured assessment endpoint from the available data. The
variance of a single predicted Y-value for a given X-value is the appropriate
value to use in calculating confidence intervals. The interest is in the
uncertainty concerning an individual future observation of Y, such as toxic
threshold, for an untested species-chemical combination. If more than one
extrapolation is required, the Y-value from the first extrapolation becomes
the X-variable in the second extrapolation.
The variance is equal to the sum of the variances from the individual
extrapolations. The variance associated with the extrapolation would depend
in part on variation in the real world, e.g., inter-species differences. The
variance would also reflect the number of measured data points available for
the extrapolation. Thus, the variance associated with the extrapolation would
be higher with less well-characterized relationships.
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Users Manual for Eco. Risk Barnthouse e_t aj,. 1982, 1986 Page C-70
Barnthouse and associates validated the double extrapolation method for
estimating an MATC of one aquatic species from the acute LC50 of another where
laboratory data permitted. They found that the predictions based on the AEE
method were more accurate than the application of generic safety factors. The
major advantage of the AEE is that the uncertainty of each estimate is
quantified.
Exposure Assessment
Use of the AEE method presupposes that an exposure assessment has already
been performed.
Risk Characterization
The AEE method defines ecological risk as the probability that the ECC
exceeds the toxicity benchmark value (TBV). Assuming that the ECC and TBV are
independent and log-normally distributed, then: Risk - Prob(logCTBV) -
log(ECC) < 0). Figure C7-1 illustrates this risk as the shaded area of
overlap of the ECC and MATC probability distributions. Thus, the AEE method
can incorporate both the uncertainty in the value of the toxicological ••
endpoint and the uncertainty in the estimate of environmental concentrations.
v
7.1.2.3 Operational Resource Requirements
The AEE method makes maximal use of available toxicological data to
estimate unmeasured toxicological benchmarks and/or to estimate the
uncertainty associated with either measured or unmeasured TBVs. If a user is
interested in a particular extrapolation for which a regression analysis has
already been performed with a reasonably up-to-date toxicological data base,
little computational effort is required to complete an AEE assessment. If a
user intends to make a novel extrapolation, an extensive literature search and
review is required, followed by a moderate computational effort.
7.1.2.4 Summary
The AEE, like the quotient method, is best suited for answering the
question "is there a high probability of concern or not". It does not
estimate the magnitude of ecological effects expected in situations in which
the EEC exceeds the toxicological benchmark.
This technique could also be used for ranking the potential ecological
risk posed by a series of chemicals. The estimates of uncertainty associated
with each TB would help identify which ratios of EEC/TB are significantly
lower or higher than others.
Compared with the quotient method, the AEE has the advantage of defining
and quantifying the terms of an extrapolation using a large amount of
toxicological data. Measurement of the variability in species' sensitivity to
toxicants and in the relationship of various toxicological benchmarks to each
other provides an estimate of the uncertainty associated with unmeasured
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coxicological benchmarks. The explicit estimation of uncertainty should
assign higher probabilities of adverse effects for chemicals which are less
we11-characterized.
7.1.3 EXTRAPOLATION OF POPULATION RESPONSES
7.1.3.1 Introduction
The purpose of the extrapolation of population responses (EPR) method is
to extrapolate from individual-level responses in laboratory settings to
population-level responses in the field. The EPR method first develops life-
stage-specific exposure-response functions for a single species. These form
the basis for estimating changes in a population's reproductive potential with
toxicant concentration as well as the confidence limits of the relationship.
7.1.3.2 Description of Method
Receptor Characterization
Application of the EPR method requires a life-stage characterization of a
receptor population. Barnthouse et al. (1986) developed an example for a fish
population in which all life stages are exposed through the same medium. The
number of life stages is variable, depending on the organism and the quality
of data available for each stage. A life table including the following
information is required for each age after the first breeding year: the
proportion of mature females, the fecundity per mature female, and the
cumulative probability of survival from the age of reproductive maturity to
each future age.
Hazard Assessment
The life-stage-specific concentration- or exposure-response functions
form the backbone of the EPR method. Exposure-response data sets can be
fitted to a logistic equation using nonlinear least squares regression. The
confidence limits (uncertainty band around the fitted regression) can be
estimated from the elements of the variance-covariance matrix. Because full
life-cycle exposure-response data are rarely available, extrapolation from the
few well studied species will often be necessary.
Barnthouse ejj a\. (1986) used the analysis of extrapolation error (AEE)
method to estimate the chronic LC25 and one of the exponential parameters for
exposure-response functions for 60 species/contaminant/experimental condition
combinations. The extrapolated response functions and the uncertainty bands
were verified for the 60 data sets with favorable results.
Exposure Assessment
This method presupposes an exposure assessment.
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Users Manual for Eco. Risk Barnthouse et a^. 1982, 1986 Page C-73
Risk Characterization
The endpoint considered for the risk assessment in the Barnthouse et al.
(1986) example was the reproductive potential of a female recruit into a fish
population. Lifetime reproductive potential is dependent on the probability
of survival to successive life stages and age-specific fecundity. From the
exposure-response curves developed for each life stage, the probability of
survival from one age to the next is estimated. Finally, the exposure -
response function for reproductive potential is constructed using a log-probit
model. The appropriate confidence limits are used to define a band of
uncertainty around the exposure-response function as illustrated in Figure
C7-2. Barnthouse et. al. (1986) found narrow confidence limits associated with
between species extrapolations and large confidence limits in situations where
chronic benchmarks were extrapolated from acute values.
The final step is estimating the probability of various effect levels
given an environmental concentration. Barnthouse e_£ al. (1986) did not
elaborate on this point. From Figure C7-2, it is clear that the most simple
approach would be to define concentration levels of no concern, intermediate
concern, or serious concern. Alternatively, one could define the probability
of exceeding a certain percentage reduction of female reproductive potential.
7.1.3.3 Operational Resource Requirements
If full life-cycle exposure-response data are not available for a species
of interest, the function could be extrapolated from information on other
species. The degree of effort would be entirely dependent on the number of
extrapolations required. Barnthouse et al. (1986) have compiled and presented
a large body of freshwater fish acute and chronic toxicity test information.
A computer software statistical package is required for fitting experimental
data to log-probit regressions. The remaining computations require a moderate
amount of computer assistance.
7.1.3.4 Summary
The extrapolation of population responses (EPR) method is very similar co
simple analysis of extrapolation error with two possible improvements. The
EPR method combines information from several life stages into a single effect
level, thus offering the potential for a more accurate representation.
Second, a variety of effect levels can be considered as endpoints.
The EPR methodology provides several advantages over the use of MATCs for
setting regulatory standards for water criteria. Barnthouse et al. have
pointed out that the use of hypothesis testing to estimate MATCs leads to
nonconservative risk assessments. The probability of committing type II error
(i.e., accepting a concentration level as non-toxic.when it is in fact toxic)
is unconstrained. The less rigorously the experiment is conducted (smaller
sample sizes or poor control of other environmental factors), the higher a
concentration will have to be to produce response levels that are
significantly different from control levels. With the EPR method, Barnthouse
e_t al. estimated LC^QS for several species that were an order of magnitude
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Users Manual for Eco. Risk
Barnthouse e£
FIGURE C7-2
1982, 1986
Page C-74
PERCENT RESPONSE VS. CONCENTRATION
(Source: Barnthouse g£ al. 1986)
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Users Manual for Eco. Risk Barnthouse §_£ aJL. 1982, 1986 Page C-75
lower than the estimated MATCs. For example, che brook trout MATC for
methylmercury (0.53 ug/liter) corresponds to a 60 to 78 percent reduction in
reproductive potential in brook trout according to the EPR method.
A limitation of this type of model, however, is that a large amount of
the toxicant-induced mortality that affects female reproductive potential
occurs in the young, pre-reproductive age-classes. If recruitment into the
breeding population in free-living populations is in part density-dependent, a
given percentage death in a pre-reproductive cohort might have a much smaller
percentage effect on recruitment into the breeding population.
7.1.4 ECOSYSTEM UNCERTAINTY ANALYSIS
7.1.4.1 Introduction
The ecosystem uncertainty analysis (EUA) method attempts to extrapolate
from single species toxicity effects to ecosystem level effects. In
particular, the model attempts to address higher order effects for
concentrations that are well below acute LC^Q values. The EUA emphasizes the
uncertainty associated with each level of extrapolation. In principle, any
sort of model of biotic community relationships could be used. The key
feature of the uncertainty analysis is the use of simulation techniques and
probability distributions for underlying parameters to generate probability
distributions for effects.
7.1.4.2 Description of Method
Receptor Characterization
Receptor characterization depends upon the actual "ecosystem" model used.
As an example, Barnthouse et al. (1986) used the Standard Water Column Model
(SWACOM). SWACOM is a computer model designed to simulate the pelagic
portions of a north temperate lake ecosystem. The model, illustrated in
Figure C7-3, includes ten phytoplankton populations, five zooplankton
populations, three planktivorous fish populations, and a top carnivore
population. The populations within each trophic level may have differing
sensitivities to toxicants. Seasonal variation in abiotic factors constrain
the seasonal behavior of the populations and their relationships.
Hazard Assessment
SWACOM model parameters include processes such as grazing, respiration,
and susceptibility to predation. The expected change in each parameter value
in response to exposure to a given toxicant concentration is expressed by an
element of an effects matrix. For each of the effects in the matrix, an
estimation of uncertainty is required. The uncertainty can either be measured
from appropriate data or estimated based on professional judgment. Usually,
however, the appropriate data are not available for the parameters used by
SWACOM.
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Users Manual for Eco. Risk Barnthouse et aj.. 1982, 1986
FIGURE C7-3
THE SWACQtf GOHFOTER MODEL
(Source: Barnthouse et al. 1986)
Page C-76
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Users Manual for Eco. Risk Barnthouse gt aJL. 1982, 1986 Page C-77
To circumvent the lack of data relating toxicant concentrations to the
parameters used by the SWACOM model, Barnthouse and associates suggested using
a "general stress syndrome" to depict each population's response to sublethal
toxicant concentrations. For example, a general stress response by grazing
fish might include decreased respiration, lower temperature optima, and
increased mortality and susceptibility to predation. To extrapolate from
acute toxicity LC5Q values to effects on respiration, susceptibility to
predation, etc., Barnthouse eg aj.. (1986) assumed that (1) organisms respond
to all toxicants in a uniform manner, and (2) all parameters (e.g.,
respiration, susceptibility to predation) are changed by the same percentage
in response to a toxicant exposure.
This extrapolation is extremely tenuous and does not lend itself to an
analysis of uncertainty. It is necessary to know not only the percentage
change in parameters, but also the uncertainty to be associated with the
change. Barnthouse e_t al. (1986) assumed that "all parameter changes have an
associated uncertainty of plus or minus 100%."
Exposure Assessment
This aspect of modeling was not addressed.
Risk Characterization
The endpoints of EUA analysis are operationally defined in terms of
"effects variables". These must be predetermined and might include effects
such as a 25% decrease in game fish biomass or a 50% increase in algal
biomass.
SWACOM simulates the population dynamics of the phytoplankton,
zooplankton, grazing fish, and one top carnivore. Competition among species
within a trophic level, and predation or grazing by higher trophic levels on
lower trophic levels, are modeled. In addition, an annual cycle of nutrient
flux, light, and temperature affect growth and reproduction. Environmental
concentrations are compared to the effects matrix for each species and each
parameter to determine the direction and magnitude of the effect.
The ecosystem is then modeled using a Monte Carlo simulation. At the
start of each run, parameter values are drawn from their statistical
distributions. In this way, uncertainties are propagated through the model
ecosystem. The simulation continues until a stable frequency distribution of
results is obtained. In this way, the uncertainties of extrapolating
laboratory data to the field should become statements about the uncertainty of
an undesirable effect. Thus, the risk of a specified adverse effect level is
a function of both of direct toxicity and the effect of uncertainty resulting
from the extrapolation.
The ecosystem effects calculated by EUA could not be predicted without
this form of analysis. For example, increasing bluegreen algal blooms with
increasing toxicant concentration appears counterintuitive. The SWACOM
adaptation, however, indicated that even though a compound may be toxic to the
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Users Manual for Eco. Risk Barnthouse e_t al. 1982, 1986 Page C-78
algae, reduction in sensitive grazing organisms can more than compensate for
the adverse direct effects on phytoplankton.
7.1.4.3 Operational Resource Requirements
In general, the data requirements for ecological uncertainty analysis are
large. First, a reasonably large number of receptor populations are required
to simulate a microcosm. Second, exposure-response data, mechanisms of
action, and sublethal toxicant effects are required for reasonably realistic
modeling efforts using a model such as SWACOM. Finally, a large number of
input parameter distributions must be obtained or estimated.
To preserve the influence of the uncertainty in the input parameter
distributions, Monte Carlo simulation techniques are most appropriate. The
number of iterations of the model required to generate a reasonably stable
output distribution can in theory be-quite high.
7.1.4.4 Summary
The compromises that are needed to implement a model like SWACOM reflect
a fundamental shortcoming in available data. Some of the more realistic
options for modeling the effects of toxicants at population and community
levels require far more information on sublethal effects than simple acute
toxicity or even chronic toxicity testing typically provide. To the extent
that models like SWACOM are scientifically verified (e.g., scientifically
reasonable), the need for qualitatively different toxicological data is
apparent. The alternative solution is to use models that use mortality as an
endpoint, as does the CERCLA Type A Assessment Model (DOI 1987a).
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Regional Ecological Assessments Ballou et al. 1981 Page C-79
7.2 REGIONAL ECOLOGICAL ASSESSMENTS: CONCEPTS, PROCEDURE AND APPLICATIONS
(Ballou e£ ai. 1981)
Argonne National Laboratory (ANL) developed several approaches for
conducting regional assessments of the potential ecological impact of energy-
development activities. Some of the primary approaches proposed include (1) a
regional ecological characterization scheme, (2) procedures for estimating
crops and natural vegetation at risk from gaseous pollutants, (3) procedures
for estimating crop yield reductions, (4) procedures for estimating and
quantifying direct land disturbance, (5) a data base and mapping procedure for
predicting overlap with endangered or threatened species habitat, and (6) the
use of ecosystem resiliency and diversity as a tool for ecological assessment.
The scope of the project focused primarily on energy impacts in the Midwestern
United States.
Below, each of these approaches is discussed in the context of its
usefulness for ecological risk assessments.
7.2.1 REGIONAL ECOLOGICAL ASSESSMENT UNITS
7.2.1.1 Introduction
ANL developed an ecological receptor characterization scheme based on
regional physiographic units. This classification scheme reflects regional
homogeneity of environmental variables, and thus is based on ecologically
meaningful characteristics. Such a classification scheme facilitates
generalizations regarding the impacts of energy-related or other activities on
regionally-homogeneous ecological areas.
7.2.1.2 Description of Method
Receptor Characterization
ANL categorized the terrestrial environment into approximately 90
"ecological assessment units (EAUs)" defined by commonalities of
bedrock/substrate, landform, soils, natural vegetation, and land use (e.g.
Piedmont lowland, unglaciated Allegheny plateau). The EAUs were built from an
existing national classification system based on climate, soils, and
topography. EAUs were derived from physiographic units that were
statistically homogeneous according to land use and cover type. Within each
of the EAUs, ANL developed a county-specific data base that included the
generalized land-use characteristics (e.g. pasture, cropland) and potential
natural vegetation (e.g. maple-basswood forest, oak savannah), and the percent
f cover for each county by these categories.
7.2.1.3 Operational Resource Requirements
,*
This approach for receptor characterization is based on readily available
information. The initial data gathering and compilation would be resource-
intensive. However, because the initial collection and compilation has been
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Regional Ecological Assessments Ballou et al. 1981 Page C-80
completed by ANL, application of this approach should be inexpensive and
require minimal individual expertise and effort.
7.2.1.4 Summary
The major advantage of the EAU approach for receptor characterization is
that large-scale energy activities or other large-scale ecological stresses
can be related to broad, ecologically similar regions. For example, the
Adirondack Mountains, Allegheny Mountains, and Superior Uplands are each
physiographic units that are highly sensitive to acid rain deposition because
they are characterized by bedrock and soils with low buffering capacities due
to a lack of base minerals. Based on similar interpretations of common
characteristics among EAUs, potential impacts to related ecosystems can be
estimated. This approach is for receptor characterization only, and must be
combined with hazard and exposure information for use in ecological risk
assessments.
7.2.2 DETERMINATION OF CROPS AT RISK AND ESTIMATING CROP YIELD REDUCTIONS
7.2.2.1 Introduction
One of the major potential impacts of fossil-fuel burning plants is
reduced crop yield and damage to both agricultural and natural vegetation from
elevated sulfur dioxide (SC^) levels. ANL used reported exposure-response
data derived from energy technology assessments to identify species of crops
and natural vegetation types at risk from future releases of S02. Also,
exposure-response data were combined with model-based estimates of exposure to
determine crop yield reductions from estimated releases. Yield reductions
were determined for both acute and chronic exposure scenarios. The risk of
significant crop yield reduction (i.e., > 5%) was determined by the number of
concentration threshold exceedances.
7.2.2.2 Description of Method
Receptor Characterization
ANL identified the dominant crop species and natural vegetation species
within each county of the U.S. Crop data by county was obtained from the 1974
Census of Agriculture. Information on natural vegetation communities was
derived vising Kuchler's (1974) classification of vegetative communities in Che
U.S. (Am. Gco. Soc. Sp. Pub. No. 36). Dominant vegetative species were
assumed to be representative of other vegetation within the community. Only
terrestrial species were considered in the classification. The most
susceptible stage of growth for each species was identified.
Hazard Assessment
The hazard assessment consists of two phases: a screening phase in which
available toxicological data are used to determine the relative sensitivities
of the various dominant vegetation types to S02, and a more detailed
assessment phase in which available toxicological data are used in a
quantitative determination of crop-specific effect levels. The screening
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Regional Ecological Assessments Ballou et al. 1981 Page C-81
phase is limited to determination of impacts following acute, high-level
exposures. The more detailed assessment phase evaluates impacts following
both acute and chronic exposures. Presumably, the most sensitive endpoint was
used in determining relative species sensitivities. Although species-
specific data are used, the results of the assessment are assumed to be
representative of potential community impacts. There is no treatment of
uncertainty in the hazard assessment.
In the screening phase, ANL classified 93 of their 116 vegetation
categories according the sensitivity of the dominant plants to S02- Plants
were classified qualitatively as sensitive, intermediate, and resistent. The
classification was based on an extensive literature review of the exposure-
response of a variety of plants to S02- Average 3 -hour maximum $©2
concentrations were used to predict impacts on vegetation.
In the detailed phase, ANL used 'estimates of short-term peak
concentrations to determine impacts on crop yield following acute exposures .
Impacts following chronic low- level exposure were determined using studies on
yield reduction in soybeans. Based on the available literature, a damage
threshold of 234 ug/nr was determined for soybeans. Although a hazard
assessment for chronic exposures was conducted for soybeans only, the approach
of threshold identification can be used for any species for which adequate
exposure -response data are available.
Exposure Assessment
Exposure concentrations were determined for both short-term and long-term
releases. Spatial and temporal variations in exposure concentrations were
considered using short-term and long-term dispersion models. Only exposure
via air was considered.
Acute exposures to SOo. ANL developed modeling techniques to estimate
the peak ground- level concentrations of air pollutants near a point source as
a function of power plant capacity and emission rate. The techniques used
short-term mathematical models to provide estimates of 'peak ground- level
concentrations and the distances from the source at which they occur. In chis
screening methodology, the critical meteorological conditions under which peak
concentrations may occur are identified and the corresponding concentrations
determined. Thus, the worst case scenarios are modeled and the ground surface
area exposed to various peak SC>2 levels is defined.
Air Quality Modeling. A slightly different air quality
modeling approach was used to estimate doses following long-term, low- level
pollutant exposure. Under this approach, the total number of multiple-hour
exposure periods in which a given threshold concentration is exceeded is
determined. For the purposes of cheir assessment of soybean yield loss, ANL
chose a threshold value of 234 ug/nr for a 4-hour exposure period based on
empirical data. A Gaussian dispersion model was used to determine exposure
concentrations. Total dose was determined by summing the total accumulated
dose for each 4-hour period in which the threshold value was exceeded.
Spatial contours of the total SC>2 dose were developed.
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Regional Ecological Assessments Ballou et al. 1981 Page C-82
Risk Characterization
The contours of the total S02 dose determined in the exposure assessment
are used to estimate the risk of significant crop yield reduction by comparing
estimated concentrations with available crop dose-response information.
Under this quantitative approach, only potential damage to a single crop
species following exposures to a single compound via a single route of
exposure are evaluated. There is no treatment of uncertainty. The approach
has been partially validated by comparing modeling results with empirical
data, and the results were in good agreement. The approach is flexible, so
that if sufficient dose-response information is available, it can be applied
to similar situations to determine potential damage from airborne pollutants
to other plant species.
7.2.2.3 Operational Resource Requirements
This approach is expensive and labor-intensive. Large amounts of data
must be compiled to determine dominant species within counties or other
designated regions, and to determine dose-response relationships for each
species and each chemical evaluated. Additional expertise and costs are
associated with the air modeling effort. The costs and level of skill
associated with interpretation of the available data are moderate.
7.2.2.4
The procedure for crop loss assessment proposed by ANL is a quantitative
worst-case approach to estimating the areal extent of significant crop yield
reduction (i.e., in excess of 5 percent) from emissions of phytotoxic gases.
The primary advantages of this approach are that it is flexible and can be
applied to estimate damage to other plant species from various air pollutants.
However, the approach is cost- and data-intensive, and its applicability is
limited somewhat by a current lack of dose-response data for many species and
pollutants. In addition, only damage co plants is estimated and is assumed to
be representative of damage to the community or ecosystem.
7.2.3 DETERMINATION OF LAND USE DISTURBANCES
7.2.3.1 Introduction
ANL developed algorithms for quantifying the amount of land disturbed by
energy activities. The proposed methodology determines the amount of land
required for power plant sites or for mining, the amount of additional land
disturbed by these activities, and the significance of the impacts. Under
this approach, the amount of forest, grassland, or other natural habitat
disturbed is of greatest concern for ecological assessments. Impacts from a
number of processes are considered, including building, moving soil, noise,
human presence, road construction, and traffic.
The endpoint of concern in this assessment is loss of land use. The
stresses in this assessment are direct and indirect physical stresses (e.g.,
land destruction (direct), human presence and noise (indirect)) rather than
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Regional Ecological Assessments Ballou et al. 1981 Page C-83
chemical, and therefore hazard assessments as typically conceived are not
appropriate. Exposure assessment determines the amount of overlap of the
physical disturbances with the land. "Risk" in the context of this approach
is not a determination of the probability of an effect level, but rather an
estimate of damage (land usage loss).
Below, we discuss the aspects of this approach that are applicable to
risk assessment.
7.2.3.2 Description of Method
Receptor Characterization
The receptor under this approach is land, and land is differentiated by
land use. Land use categories include crop land, pasture, range, forest,
urban land, Federally-owned land, and-other rural land (e.g., roads, farm
buildings, swamps, and barren land).' It was assumed that Federal land is
covered by natural vegetation, and that very little urban land remains in its
natural state. Land use was determined on a county basis, and is therefore
relatively site-specific.
Exposure Assessment •.'
As mentioned above, exposure assessment is limited to a determination of
the overlap of physical disturbances with land. Areas of directly disturbed
land are determined on the basis of a step function using average values
reported in the literature for particular technologies and generating
capacities. In this way, basic site requirements for power plants and mines
are determined.
Risk Characterization
Two approaches are used. The first approach is quantitative and equates
damage with land loss. The second is qualitative and attempts to interpret
the effects of land loss on the affected species. Under this qualitative
approach, species expected to occur in a particular land type are identified,
and the habitat requirements for these species are outlined. Land use
disturbance is interpreted in the context of loss of habitat requirements.
7.2.3.3 Operational Resource Requirements
The initial classification of land use is moderately labor- and cost-
intensive. Determination of airect land loss requires minimal resources and
expertise. Qualitative assessments of potential impacts to nonplant species
requires personnel trained in wildlife biology.
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Regional Ecological Assessments Ballou et al, 1981 Page C-84
7.2.3.4 Summary
This approach is unique among many of the assessment methodologies in
that land loss and disturbance are endpoints of concern. The approach is
limited to interpretation of effects associated with physical disturbances of
land use; procedures for predicting and quantifying land use loss from
chemical contamination have not yet been developed.
7.2.4 POTENTIAL IMPACTS TO ENDANGERED AND THREATENED SPECIES
7.2.4.1 Introduction
ANL documented the distribution of endangered and threatened species and
their habitats in the Midwest and the-reasons for the declining populations.
The approach is not an ecological risk assessment in that no hazard, exposure,
or risk assessments are conducted. The approach does, however, provide some
useful concepts for receptor characterization; these are discussed below.
7.2.4.2 Description of Method
Receptor Characterization
ANL conducted an in-depth literature search and consulted state and
Federal wildlife experts regarding the distribution of threatened and
endangered species in the Midwest. Particular emphasis was placed on habicat
characteristics and habitat requirements in the context of interpreting
species decline. Such information is useful in predicting the potential
response of a species to environmental changes.
The county-level distribution of designated species was determined. The
resulting data base was designed for use as a screening mechanism to identify
counties where potential conflicts may occur through the implementation of
projected energy scenarios.
7.2.4.3 Operational Resource Requirements
This approach for receptor characterization is based on readily available
information. The initial data collection and compilation is moderately
resource-intensive and requires some expertise in the area of ecology.
However, once the data are compiled, application of the approach should be
inexpensive. Interpretation of potential impacts from environmental
disturbances requires some understanding of ecology.
7.2.4.4 Summary
This approach is an easily-implemented, relatively low-cost, qualitative
screening mechanism for evaluating the potential for damage Co endangered and
threatened species. It does not predict what effects, if any, will occur, but
only estimates the potential for effects. Counties with more endangered
species can be assumed to have a greater potential for detrimental effects.
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Regional Ecological Assessments Ballou et al. 1981 Page C-85
No hazard, exposure, or risk components are incorporated into this approach;
it is mainly applicable for receptor characterization and as a screening tool.
7.2.5 SPECIES DIVERSITY AS A TOOL IN ECOLOGICAL ASSESSMENTS
7.2.5.1 Introduction
ANL proposed a conceptual approach for incorporating ecosystem diversity
measurements in ecological field assessments. Species diversity is an
attribute of all ecosystems, and is often considered a primary indicator of
ecosystem "health", stability, and resilience. Ecosystem resilience is a
system's ability to recover without losing its intrinsic identity. The more
resilient a system, the better that system is able to recover from
environmental perturbations. Diversity indices can be used as a measure of
ecosystem resilience. Several diversity indices were discussed.
The diversity approach is a tool for evaluating and monitoring the
ecological environment. The approach provides unique concepts for receptor
characterization, and can be used as a rough screening mechanism for
estimating ecosystem vulnerability.
7.2.5.2 Description of Method
Jlhj^acterigation
The receptors under this approach are communities and ecosystems.
Communities and ecosystems are defined both temporally and spatially, in
addition to taxonomically . Abiotic and biotic characteristics of the system
are defined and considered when determining diversity. Three types of
diversity are defined: alpha diversity refers to species diversity at a
particular site; beta diversity refers to species diversity among ecologically
similar sites; and gamma diversity refers to species diversity among
ecologically similar and variable sites. The approach results in a fairly
complete characterization of communities and ecosystems.
7.2.5.3 Operational Resource Requirements
This approach to receptor characterization is resource- and data-
intensive. Large amounts of data are needed on species composition, number,
and distributions within and between systems. This requires extensive review
of available data, and probably requires additional data gathering in the
field. Experienced ecologists would be needed to collect and interpret the
data.
7.2.5.4 Summary
This approach characterizes receptors at the community and ecosystem
level. Abiotic and biotic as well as temporal and spatial aspects of systems
are considered. The approach is not a risk assessment methodology; it is an
environmental monitoring and evaluation technique. However, the diversity of
a system can be used as a relative measure of the potential for damage
following environmental disturbances. In this context, the approach is most.
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Regional Ecological Assessments Ballou et al. 1981 Page C-86
appropriately used as a screening tool in ecological assessments. The
approach, however, is relatively resource-intensive, and this may limit its
usefulness.
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Computer Simulation Model Eschenroeder et al. 1980 Page C-87
7.3 COMPUTER. SIMULATION MODELS FOR ASSESSMENT OF TOXIC SUBSTANCES
(Eachanroeder e£ al. 1980)
7.3.1 Introduction
The computer simulation model described in this document is the result of
an exploratory study, funded by the National Science Foundation, to develop a
model to aid in the assessment of environmental effects of toxic substances.
The model is a series of generalized differential equations based on
bioenergetic parameters. The equations are used to calculate population
biomass, expressed as biomass/energy equivalents, and concentrations of the
chemical in each population. Lethal and sublethal effects of chemicals can
also be incorporated into the model and the subsequent effects on biomass and
chemical concentration in each population examined. A simplified example, DDT
in a pond ecosystem, was described in-the document. In the example, only
aquatic organisms were modeled, and populations were grouped by trophic level.
7.3.2 Description of Method
Receptor Characterization
Although the model could be adapted to simulate many systems, it is
applied in this document to a simple aquatic community. The community
consists of one or two species at each trophic level. Populations of species
at each trophic level are "lumped" and changes are simulated by each trophic
level. Species that change trophic levels in their lifetimes (e.g., perch,
which become pisciverous with age) are divided, based on biomass, into
different trophic levels. Although temporal changes are not incorporated into
this example, the possibility of incorporating it in future versions of the
model is discussed.
Hazard Assessment
Damage functions/expressed in terras of dose, were constructed for rates
of mortality, respiration, and feeding. The functions were constructed based
on assumed baseline and saturation levels for the effect of concern, and a
threshold concentration of contaminant eliciting the effect. Individuals were
then assumed to be normally distributed with respect to the step-function.
Exposure Assessment
Exposure is calculated for each compartment, in this case, by trophic
level. Chemical flux into the producer compartment is expressed as mass units
of toxic material per kilocalorie of biomass/energy. Loss terms from the
producer compartment include predation, mortality, and excretion. At higher
trophic levels, exposure occurs only through the food chain. Thus, the
chemical concentration in higher trophic levels is calculated as the influx by
feeding minus loss terms including predation, mortality, and excretion. The
incorporation of a bioconcentratton factor from water into higher trophic
levels was discussed as a possibility for future versions of the model.
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Computer Simulation Model Eschenroeder et al. 1980 Page C-88
Risk Characterization
The model uses two sets of differential equations to describe population
dynamics and chemical concentration. In the example, the model was used to
estimate change in biomass and chemical concentration in a simple aquatic
system by trophic level. Toxic effects of chemicals are combined with
exposure (i.e., dose) in the model by constructing injury functions. In this
way, effects of chemical concentrations are fed back into the set of
differential equations describing population dynamics. Risk is then expressed
as the percent reduction in energy/biomass for each compartment. In the
simplified version of the model, species were aggregated and competition was
ignored, although with sufficient information, these two issues could be
incorporated.
Predictions of residues were roughly validated using general information
on DDT levels in different trophic levels of the environment compiled from the
literature. The model was analyzed for sensitivity to effects at different
trophic levels, and to variations in the damage functions. The model could be
validated using a simple pond mesocosm or microcosm.
7.3.3 Operational Resource Requirements
The model requires some exposure-response data that are largely
unavailable (e.g., for ingestion) for endpoints that might be difficult to
measure for many systems, (e.g., bioraass, respiration rates, feeding rates).
It requires a high level of effort and skill, as well as a computer.
7.3.4 Summary
A major strength of the model is that it is very flexible. Although the
compartments in the example were trophic levels, with sufficient information,
compartments could be species or even age or weight class within species.
Stochastic and temporal elements could also be incorporated. Sublethal
effects such as feeding rate changes, as well as mortality, are included.
Risk is presented as a reduction in bioraass, which could easily be input to
costing models. Because biomass and residue concentration were modeled on a
time-specific basis, recovery dynamics of the system can be examined. A
limitation of the model is that it uses bioenergetic terms and units that are
unfamiliar to many people. Data needs are high, and with refinement of the
model, data needs would be even greater. For this reason, the model is
presently limited to very simple systems.
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Comparative Risk Project EPA/OPPE 1987 Page C-89
7.4 UNFINISHED BUSINESS: A COMPARATIVE ASSESSMENT OF ENVIRONMENTAL PROBLEMS.
APPENDIX III, ECOLOGICAL RISK WORK GROUP (EPA/OPPE 1987)
7.4.1 Introduction
The objective of the Comparative Risk Project (CRP) was to estimate and
rank current cancer risks, noncancer health risks, welfare effects, and
ecological effects presented by 31 major environmental problem areas which EPA
has some responsibility and authority to control; the purpose was to establish
program and budget priorities. Separate workgroups were formed to evaluate
each type of risk; this summary addresses only the ecological risks component
evaluated by the Ecological Risk Work Group (ERVG), which was assisted by an
expert panel convened by the Cornell Ecosystems Research Center.
The objective of the ERWG was to rank the relative ecological risks,
under existing levels of control, posed by widely disparate categories of
environmental problem areas including accidental oil spills, active hazardous
waste sites, and depletion of the stratospheric ozone layer. Because of
inconsistencies in the scope and scale of each problem area, types of
potential and actual ecological damage caused by each problem area, and levels
of data availability for each problem area, the ERWG recognized the need for a
methodology that differed from a formal risk assessment procedure.
Consequently, the ERWG used a semi-quantitative damage assessment procedure
that relied in large part on the considered judgment of experts.
The procedure followed by the ERWG and the expert panel involved five
major steps. First, the specific types of ecological stress agents associated
with each environmental problem area were identified. Second, a set of
terrestrial and aquatic ecosystem categories was defined. In the third step,
the expert panel qualitatively evaluated the potential effects of each stress
agent on the structure and function of each type of ecosystem, the
reversibility of these impacts, and the time it would take for the ecosystem
to recover after the stress agent was removed, and characterized the
geographic scale of these effects. In the fourth step, papers were prepared
for each problem area summarizing available information on sources and
emissions, exposure, potential impacts on ecosystems, level of control, and
quality of the available information. Finally, using these papers, each
member of the ERWG subjectively prepared an aggregate ranking of each problem
area (low, medium, or high) based on a separate ranking of effects on each
ecosystem. The individual rankings were tabulated, and a final ranking was
determined by consensus.
7.4.2 Description of Method
Receptor Characterization
The analysis focused on a generic set of 16 different freshwater, marine,
estuarine, terrestrial, and wetland ecosystems. Analysis of hazard included
consideration of impacts on community structure (e.g., alterations in trophic
structure or species diversity), community function (e.g., alterations in
primary production or rates of nutrient cycling), and particular species
(e.g., endangered or economically important species). No specific temporal ou
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Comparative Risk Project EPA/OPPE 1987 Page C-90
niche characteristics were included routinely, although they may have been
considered in some assessments.
Hazard Assessment
Hazard assessment focused mainly on a qualitative evaluation of the
potential impacts resulting from each stress agent (not from problem areas
directly). Stress agents included those transported through the atmosphere
(e.g., gaseous phytotoxicants, acid deposition), those transported through
surface water (e.g., pesticides, toxic inorganics, nutrients), those applied
directly to terrestrial systems (e.g., pesticides, solid matter), and others
(e.g., habitat alteration, ground-water contamination). Potential effects
were estimated first with respect to the scale at which impacts would likely
occur (e.g., local, regional, global). Next, the potential effects of each
stress agent on each ecosystem were -evaluated with respect to their potential
intensity (high, medium, low, no effect), the nature of the effect (e.g.,
effects on ecosystem structure and function or on endangered or economically
important species), the reversibility of the impacts, and the probable time
scale for recovery following removal of the stress agent (years, decades,
centuries, indefinite).
In most cases, the assessment was based on known effects on surrogate
individual/population receptors rather than on direct ecosystem dose-response
data. For some problem areas with a small number of stress agents (e.g., acid
deposition), data on acute or chronic toxicity to individual species were used
to evaluate hazard. However, most chemical stress agents considered were
defined on the basis of chemical class (e.g., pesticides and herbicides,
nutrients, toxic organics), so it is unclear whether toxicity was assessed on
the basis of the most toxic constituent, the least toxic constituent, or some
average toxicity value.
Certain modifying factors were considered for each stress agent. Several
ecosystems (freshwater lakes, streams, and wetlands) were defined on the basis
of whether or not they were buffered. Wetlands also were defined on the basis
of whether or not they were isolated from other flowing surface water. Other
modifying factors considered in evaluating hazard included water hardness (for
inorganic chemicals) and bioaccumulation (for organic chemicals).
A qualitative assessment of the uncertainty associated with each stress
agent/ecosystem hazard assessment was included. Situations in which the data
and understanding were sufficient for certain or probable assessments were
differentiated from those in which the stress-response relationship was poorly
understood or adverse responses occurred infrequently. Qualitative
assessments of the quality of data available also were included in each
problem area summary paper.
In the final ranking scheme, the greatest weight was placed on the scale
of impact of each problem area, followed by the reversibility of potential
effects. However, no particular weighting criteria were presented. In
particular, it is unclear how the aggregate rankings of each problem area were
determined from the separate rankings of effects on each ecosystem.
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Comparative Risk Project EPA/OPPE 1987 Page C-91
Exposure Assessment
Exposure assessment was presented only in each problem area summary paper
(exposure was not considered in evaluating impacts of stress agents on
ecosystems) and included geographical extent, intensity, and frequency. All
exposure pathways relevant to a particular problem area were considered.
Exposure data consisted primarily of summary statistics on release quantities
or proportion of a given ecosystem in which contaminant concentrations were in
excess of applicable criteria. No attempt was made to provide concentration
estimation methods. Data on toxicity modifying factors (e.g., water hardness,
pH) were presented where available. Qualitative estimates of uncertainty were
included.
Risk Characterization
The approach was highly qualitative and relied heavily on expert
judgment. Maximum emphasis was placed on the integration of information on the
hazards associated with multiple contaminants, multiple routes of exposure,
and multiple stress agents associated with a particular problem area. Because
it is highly qualitative, this approach can be broadly applied to a wide range
of environmental problem areas.
Treatment of uncertainty was also highly qualitative, reflecting
uncertainty in measurements as well as biotic and ecological impacts. No
quantitative estimates (e.g., probability distributions, confidence intervals)
were possible with 'this approach.
The approach could be validated internally by having a number of
different work groups evaluate each problem area and compare their final
relative rankings. External validation, calibration, or sensitivity
evaluation would not be possible.
The approach as documented cannot be verified, because several critical
steps cannot be reconstructed. Particularly absent are algorithms for
aggregating component scores.
7.4.3 Operational Resource Requirements
For most problem areas, a limited amount of data (e.g., receptors,
chemicals) are available. The approach allows comparison of problem areas
with widely varied data bases. The cost and level of effort involved in using
this approach are low relative to those associated with a similar effort based
on quantitative risk assessments. However, proper use of this method requires
the highest level of skill, experience, and ecological knowledge.
7.4.4 Summary
The Comparative Ecological Risk approach is not a formal risk assessment
procedure aimed at quantifying ecological risks. Instead, it is a semi-
quantitative damage assessment method aimed at ranking relative ecological
risks posed by major environmental problem areas. Such a method is probably
the only practical means for comparing the relative risks associated with
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Comparative Risk Project EPA/OPPE 1987 Page C-92
widely disparate problem areas, particularly on a regional or national level.
Its strengths lie in its ability to provide a broad integration of data on
known exposures, potential hazards, and multiple pathways and stresses. For
example, the problem areas receiving the highest risk ranking (stratospheric
ozone depletion, CC>2 and global warming) were not those for which data were
most available. The main limitations of this approach are its high reliance
on qualitative judgments and the absence of repeatability due to a lack of
sufficient documentation for some of the critical steps.
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