ECOLOGICAL RISK ASSESSMENT
GUIDANCE FOR SUPERFUND:
PROCESS FOR DESIGNING AND CONDUCTING
ECOLOGICAL RISK ASSESSMENTS
INTERIM FINAL
U.S. Environmental Protection Agency
Environmental Response Team
Edison, NJ
June 5,1997
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UNITED STATES ENVIRONMENTAL PROTECTION AGENCY
WASHINGTON, D,C. 20460'
JUN -2 1397
OFFICE OF
SOLID WASTE AND EMERGENCY RESPONSE
MEMORANDUM
SUBJECT:
FROM:
TO:
Ecological Risk Assessment Guidance for Superfund, Process for Designing and
Conducting Ecological Risk Assessments (EPA 540-R-97-006)
Stephen D. Luftig, Director^
Office of Emergency and Remedial Response
Director, Office of Site Remediation and Restoration
Region I
Director, Emergency and Remedial Response Division
Region n
Director, Hazardous Waste Management Division
Regions HI, EX
Director, Waste Management Division
Region IV
Director, Superfund Division
Regions V, VI, VTI
Assistant Regional Administrator, Ecosystem Protection and Remediation
Region VTU
Director, Environmental Cleanup Office
Region X
This memorandum transmits the interim final Ecological Risk Assessment Guidance for
Superfund, Process for Designing and Conducting Ecological Risk Assessments. This guidance
was prepared to address the questions posed by Remedial Project Managers and On-Scene
Coordinators related to conducting ecological risk assessments. This guidance builds on
documents in preparation by the Office of Research and Development.
Ecological risk assessment is an important part of both the removal and remedial work
conducted at Superfund sites. The Ecological Risk Assessment Guidance for Superfund, Process
for Designing and Conducting Ecological Risk Assessments is an important step in making risk
assessments more consistent across the Regions. The guidance supersedes the Risk Assessment
Guidance for Superfund: Volume H - Environmental Evaluation Manual, 1989 (EPA/540/1-89-
Printed on Recycled Paoer
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001). The new guidance will assist the Environmental Protection Agency (EPA) in continuing to
meeting the requirements of sections 121(b)(l) and (d) of the Comprehensive Environmental
Response, Compensation, and Liability Act (CERCLA), as amended by The Superfund
Amendments and Reauthorization Act (SARA). Sections 121 (b)(0 and (d) require that remedial
actions be protective of the environment. This guidance will also assist the EPA in complying with
Section 121(c) which requires future reviews to ensure that the environment is protected at sites
where contaminants remain after remedial actions were completed.
The guidance has been extensively coordinated with the National Oceanic and
Atmospheric Administration; Department of Interior, specifically the Fish and Wildlife Service;
Department of Defense; and the Department of Energy which should. Staffs in several of the
Regions are already utilizing a draft of the guidance which should improve coordination with
other Federal Agencies and Departments and lead to more consistent and transparent risk
assessments. This guidance should be used for all ecological risk assessments conducted under
CERCLA
Scientific/Management Decision Points (SMDP) instituted in the document will bring risk
managers and Natural Resource Trustees into the risk assessment process earlier and streamline
the process. Section 104(b) (2) of CERCLA requires that the EPA promptly notify Trustees of
potential natural resource injuries and that the EPA seek to coordinate the assessments
investigations, and planning of response activities with them. As a matter of policy, the EPA
should not only comply with the statutory directives, but should also make every effort to ensure
Trustee participation at all stages of response. The SMDP are an early opportunity to coordinate
and engage the Trustees in the ecological risk assessment. These SMDP are also an opportunity
for the EPA to comply with the National Contingency Plan (40 CFR Part 300); Section 300.430
(b) (7) requires that the "EPA seek to' coordinate necessary assessments, evaluations,
investigations and planning with ...Trustees".
If you have any questions regarding the Ecological Risk Assessment Guidance for
Superfund, please contact David W. Charters, Ph.D. at (732) 906-6825 or Mark D. Sprenger,
Ph.D. at (732) 906-6826.
Attachment
cc: Department of Defense
Department of the Army
Department of the Navy
Department of the Air Force
Department of Interior
Department of Commerce, NOAA
Department of Energy
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DISCLAIMER
The policies and procedures set forth here are intended as guidance to Agency and other
government employees. They do not constitute rulemaking by the Agency, and may not be
relied on to create a substantive or procedural right enforceable by any other person. The
Government may take action that is at variance with the policies and procedures in this
manual.
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ACKNOWLEDGEMENTS
The authors wish to acknowledge all the reviewers that have assisted the authors with
insightful comments and assistance. We also wish to acknowledge the assistance of the
Response Engineering and Analytic Contract Task Leader, Mark Huston and the editorial
assistance of the ICF Consulting Group, primary editor Dr. Margaret McVey and Charles
Chappell and Kimberly Osbom.
Mark D. Sprenger, Ph.D. David W. Charters, Ph.D.
Environmental Response Team Center Environmental Response Team Center
Office of Emergency & Remedial Response Office of Emergency & Remedial Response
Primary Reviewers:
Region I
Region II
Region HI
Susan Svirsky
Patti Tyler
Shari Stevens
Barbara O Kom
Robert Davis
Region IV Lynn Wellman
Region V Brenda Jones
James Chapman, Ph.D.
Region VI Susan Roddy
Jon Rauscher, Ph.D.
Region VII Steve Wharton
Robert Koke
Region VIII Gerry Henningsen, Ph.D., D.V.M.
Dale Hoff, Ph.D.
Mark Wickstrom, D.V.M.
Region DC Clarence Callahan, Ph.D.
Ned Black, Ph.D.
Region X P. Bruce Duncan, Ph.D.
Julius Nwosu
Joe Goulet, Ph.D.
Headquarters: Steve Ells
State of Texas: Larry Champagne
U.S. Fish & Wildlife Service: Nancy Finley
Peer Review Committee:
David Anderson, Ecology & Environment, Taylor, MI
John Bascietto, DOE
Tom Campbell, Woodward Clyde, Denver, CO
Cherri Bassinger-Daniel, State of MO, Department of Health
HI
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Tom Dillon, U.S. Corps of Engineers
Alyce Fritz, NOAA
Duncan Gilroy, State of Oregon DEQ, Portland, OR
Joe Greene, U.S. EPA
Mark Harkins, Science & Space Technical Committee, Washington, DC
Chris Ingersoll, U.S. DOI/NBS, Columbia, MO
Mark Johnson, U.S. Army, Aberdeen, MD
Lawrence Kapustka, EPT, Seattle, WA
Alan Mclntosh, University of Vermont
Gary Mangels, American Cyanamid
Mary Malta, NOAA
Jennifer Roberts, DOEC, State of Alaska, Department of Environmental Conservation
Glen W. Suter, n, Martin Marietta Energy Systems, Inc., Oak Ridge National
Laboratory
Randy Wentsel, U.S. Army
Janet Whaley, U.S. Army, Aberdeen, MD
Stakeholder Meeting Attendees:
Jeff Foran, Meeting Facilitator
Judith Bland, Merck
Jim Clark, Exxon Biomedical Sciences
David Cragin, Elf Atochem
Steve Geiger, Remediation Technology
Simeon Hahn, U.S. Navy
David Hohreiter, Blasland, Bouck, and Lee
Kenneth Jenkins, consultant (Jenkins, Sanders, & Associates) representing General
Electric
Lorraine Keller, Rohm and Haas
Bryce Landenberger, Dow Chemical
Dale Marino, Eastman Kodak
Ellen Mihaich, Rhone-Poulenc
Ron Porter, U.S. Air Force
Mark Powell, Center for Risk Management at Resources for the Future
Lee Salamone, Chemical Manufacturers Association
Anne Sergeant, U.S. EPA
Jean Snider, NOAA
Ralph Stahl, DuPont
Randy Wentsel, U.S. Army
Observers at the Stakeholder Meeting:
Adam Ayers, Geraghty and Miller
Steve Ells, U.S. EPA
IV
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Paul Hirsh, Chemical Manufacturers Association
Teresa Larson, National Association of Manufacturers
Reo Menning, American Industrial Health Council
Kevin Reinert, Rohm and Haas
Phil Sandine, Environmental Liability Management
Wendy Sherman, Chemical Manufacturers Association
Todd Slater, Elf Atochem
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VI
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CONTENTS
DISCLAIMER i
ACKNOWLEDGEMENTS . . . iii
LISTS OF EXHIBITS, EXAMPLES, AND HIGHLIGHTS xi
LIST OF ACRONYMS AND ABBREVIATIONS . . xiii
PREFACE , xv
INTRODUCTION: ECOLOGICAL RISK ASSESSMENT FOR SUPERFUND 1-1
PURPOSE 1-1
SCOPE 1-1
BACKGROUND 1-1
DEFINITION OF ECOLOGICAL RISK ASSESSMENT 1-3
THE ECOLOGICAL RISK ASSESSMENT PROCESS 1-3
STEP 1: SCREENING-LEVEL PROBLEM FORMULATION AND ECOLOGICAL
EFFECTS EVALUATION 1-1
1.1 INTRODUCTION 1,1
1.2 SCREENING-LEVEL PROBLEM FORMULATION 1-1
1.2.1 Environmental Setting and Contaminants at the Site 1-2
1.2.2 Contaminant Fate and Transport 1-4
1.2.3 Epotoxicity and Potential Receptors 1-4
1.2.4 Complete Exposure Pathways 1-5
1.2.5 Assessment and Measurement Endpoints 1-7
1.3 SCREENING-LEVEL ECOLOGICAL EFFECTS EVALUATION 1-8
1.3.1 Preferred Toxicity Data 1-9
1.3.2 Dose Conversions 1-12
1.3.3 Uncertainty Assessment 1-12
1.4 SUMMARY 1-12
STEP 2: SCREENING-LEVEL EXPOSURE ESTIMATE
AND RISK CALCULATION 2-1
2.1 INTRODUCTION 2-1
2.2 SCREENING-LEVEL EXPOSURE ESTIMATES 2-1
2.2.1 Exposure Parameters 2-2
2.2.2 Uncertainty Assessment 2-3
2.3 SCREENING-LEVEL RISK CALCULATION 2-4
2.4 SCIENTIFIC/MANAGEMENT DECISION POINT (SMDP) 2-5
2.5 SUMMARY 2-6
VII
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STEP 3: BASELINE RISK ASSESSMENT PROBLEM FORMULATION 3-1
3.1 THE PROBLEM-FORMULATION PROCESS 3-1
3.2 REFINEMENT OF PRELIMINARY CONTAMINANTS OF
CONCERN 3-3
3.3 LITERATURE SEARCH ON KNOWN ECOLOGICAL EFFECTS ..... 3-4
3.4 CONTAMINANT FATE AND TRANSPORT, ECOSYSTEMS
POTENTIALLY AT RISK, AND COMPLETE EXPOSURE
PATHWAYS 3-4
3.4.1 Contaminant Fate and Transport 3-5
3.4.2 Ecosystems Potentially at Risk 3-6
3.4.3 Complete Exposure Pathways 3-7
3.5 SELECTION OF ASSESSMENT ENDPOINTS . 3-8
3.6 THE CONCEPTUAL MODEL AND RISK QUESTIONS 3-10
3.6.1 Conceptual Model 3-10
3.6.2 Risk Questions 3-14
3.7 SCIENTIFIC/MANAGEMENT DECISION POINT (SMDP) 3-15
3.8 SUMMARY 3-15
STEP 4: STUDY DESIGN AND DATA QUALITY OBJECTIVE PROCESS 4-1
4.1 ESTABLISHING MEASUREMENT ENDPOINTS 4-2
4.1.1 Species/Community/Habitat Considerations 4-5
4.1.2 Relationship of the Measurement Endpoints to the
Contaminant of Concern 4-5
4.1.3 Mechanisms of Ecoxicity 4-7
4.2 STUDY DESIGN .4-7
4.2.1 Bioaccumulation and Field Tissue Residue Studies 4-8
4.2.2 Population/Community Evaluations 4-12
4.2.3 Toxicity Testing 4-13
4.3 DATA QUALITY OBJECTIVES AND STATISTICAL
CONSIDERATIONS 4-14
4.3.1 Data Quality Objectives 4-14
4.3.2 Statistical Considerations 4-15
4.4 CONTENTS OF WORK PLAN AND SAMPLING AND ANALYSIS
PLAN , 4-15
4.4.1 Work Plan 4-16
4.4.2 Sampling and Analysis Plan . . ; • • • • 4-16
4.4.3 Field Verification of Sampling Plan and Contingency Plans 4-18
4.5 SCIENTIFIC/MANAGEMENT DECISION POINT (SMDP) 4-18
4.6 SUMMARY 4-18
STEP 5: FIELD VERIFICATION OF SAMPLING DESIGN 5-1
5.1 PURPOSE 5-1
5.2 DETERMINING SAMPLING FEASIBILITY 5-2
5.3 SCIENTIFIC/MANAGEMENT DECISION POINT (SMDP) . 5-3
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5.4 SUMMARY ; . 5-4
STEP 6: SITE INVESTIGATION . . 6-1
6.1 INTRODUCTION 6-1
6.2 SITE INVESTIGATION 6-1
6.2.1 Changing Field Conditions 6-2
6.2.2 Unexpected Nature or Extent of Contamination 6-2
6.3 ANALYSIS OF ECOLOGICAL EXPOSURES AND EFFECTS 6-3
6.3.1 Characterizing Exposures 6-3
6.3.2 Characterizing Ecological Effects 6-5
6.4 SCIENTIFIC/MANAGEMENT DECISION POINT (SMDP) .......... 6-6
6.5 SUMMARY 6-7
STEP 7: RISK CHARACTERIZATION '.' 7-1
7.1 INTRODUCTION . 7-1
7.2 RISK ESTIMATION 7-1
7.3 RISK DESCRIPTION 7-4
7.3.1 Threshold for Effects on Assessment Endpoints 7-4
7.3.2 Likelihood of Risk . . .". . . 7-5
7.3.3 Additional Risk Information : 7-5
7.4 UNCERTAINTY ANALYSIS - 7-5
7.4.1 Categories of Uncertainty 7-6
7.4.2 Tracking Uncertainties 7-7
7.5 SUMMARY . 7-7
STEP 8: RISK MANAGEMENT . . 8-1
8.1 INTRODUCTION 8-1
8.2 ECOLOGICAL RISK MANAGEMENT IN SUPERFUND 8-1
8.2.1 Other Risk Management Considerations 8-2
8.2.2 Ecological Impacts of Remedial Options 8-3
8.2.3 Monitoring 8-3
8,3 SCIENTIFIC/MANAGEMENT DECISION POINT (SMDP) 8-4
8.4 SUMMARY ,8-4
BIBLIOGRAPHY Bibliography-1
GLOSSARY Glossary-1
APPENDIX A: EXAMPLE ECOLOGICAL RISK ASSESSMENTS FOR
HYPOTHETICAL. SITES
Example 1: Copper Site A-l
Example 2: Stream DDT Site A-8
Example 3: PCB Site . A-14
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APPENDIX B: REPRESENTATIVE SAMPLING GUIDANCE DOCUMENT,
VOLUMES: ECOLOGICAL, DRAFT
U.S. Environmental Protection Agency (U.S. EPA). 1997. Representative Sampling
Guidance Document, Volume 3: Ecological, Draft. Edison, NJ: Environmental
Response Team, Office of Emergency and Remedial Response.
APPENDIX C: SUPPLEMENTAL GUIDANCE ON LITERATURE SEARCH
APPENDIX D: STATISTICAL CONSIDERATIONS
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LISTS OF EXHIBITS, EXAMPLES, AND HIGHLIGHTS
List of Exhibits
EXHIBIT 1-1: Ecological Risk Assessment Framework . . . . , 1-5
EXHIBIT 1-2: Eight-step Ecological Risk Assessment
Process for Superfund f ...... 1-9
EXHIBIT 1-3: Steps in the Ecological Risk Assessment
Process and Corresponding Decision Points in the
Superfund Process 1-10
EXHIBIT 1-4: Ecological Risk Assessment Deliverables
for the Risk Manager Ml
EXHIBIT 1-5: Ecological Risk Assessment in the Remedial
Investigation/Feasibility Study (RI/FS) Process , 1-13
EXHIBIT 1-1: List of Sensitive Environments
in the Hazard Ranking System . 1-6
EXHIBIT 6-1: Analysis Phase 6-4
EXHIBIT 7-1: Risk Characterization 7-2
EXHIBIT A-l:" Conceptual Model for the Copper Site A-5
EXHIBIT A-2: Conceptual Model for the Stream DDT Site A-l 1
EXHIBIT A-3: Conceptual Model for the Terrestrial PCB Site A-17
List of Examples
EXAMPLE 1-1: Ecotoxicity-PCB Site . 1-5
EXAMPLE 1-2: Complete Exposure Pathways for Mammals-PCB Site 1-8
EXAMPLE 3-1: Exposure Pathway Model-DDT Site -...-. 3-7
EXAMPLE 3-2: Potential for Food Chain Transfer-Copper
and DDT Sites 3-8
EXAMPLE 3-3: Assessment Endpoint Selection-DDT,
Copper, and PCB Sites 3-11
EXAMPLE 3-4: Description of the Conceptual Model-DDT Site 3-12
EXAMPLE 3-5: Conceptual Model Diagram-DDT Site 3-13
EXAMPLE 4-1: Lines of Evidence-Copper Site 4-4
EXAMPLE 4-2: Selecting Measurement Endpoints-DDT Site 4-6
EXAMPLE 4-3: Tissue Residue Studies-DDT Site 4-9
EXAMPLE 5-1: Field Verification of Sampling Design-Copper Site 5-4
EXAMPLE 5-2: Field Verification of Sampling Design-DDT Site 5-5
EXAMPLE 6-1: Fish Sampling Contingency Plan-DDT Site 6-2
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List of Highlights
HIGHLIGHT 1-1: The RI/FS Process 1-2
HIGHLIGHT 1-2: Example Assessment Endpoints 1-6
HIGHLIGHT 1-3: Example Measurement Endpoints . . 1-6
HIGHLIGHT 1-4: Ecological Impact and Risk Assessment 1-8
HIGHLIGHT 1-1: Screening-level Risk Assessments 1-2
HIGHLIGHT 1-2: Industrial or Urban Settings 1-4
HIGHLIGHT 1-3: Exposure Pathway and Exposure Route 1-7
HIGHLIGHT 1-4: Non-Chemical Stressors 1-9
HIGHLIGHT 1-5: Data Hierarchy for Deriving Screening
Ecotoxiciry Values 1-10
NOAEL Preferred to LOAEL 1-11
Area Use Factor 2-2
Hazard Index (HI) Calculation 2-5
Tiering an Ecological Risk Assessment . 3-3
Environmental Fate and Exposure 3-5
Definitions: Null and Test Hypotheses 3-14'
Importance of Distinguishing Measurement
from Assessment Endpoints . 4-3
HIGHLIGHT 4-2: Terminology and Definitions 4-6
HIGHLIGHT 4-3: Elements of a QAPP 4-17
HIGHLIGHT 6-1: Uncertainty in Exposure Models 6-5
HIGHLIGHT 1-6:
HIGHLIGHT 2-1:
HIGHLIGHT 2-2:
HIGHLIGHT 3-1:
HIGHLIGHT 3-2:
HIGHLIGHT 3-3:
HIGHLIGHT 4-1:
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LIST OF ACRONYMS AND ABBREVIATIONS
AQUIRE: U.S. EPA's AQUatic Information REtrieval database
ARAR: Applicable or Relevant and Appropriate Requirements
ASTM: American Society of Testing and Materials
BAF: Bioaccumulation Factor
BCF: . Bioconcentration Factor
BIOSIS: Biosciences Information Services
BTAG: Biological Technical Assistance Group
CERCLA: Comprehensive Environmental Response, Compensation, and Liability Act
CLP: Contract Laboratory Program
DDT: Dichlorodiphenyltrichloroethane
DQO: Data Quality Objective
EC50: Effective Concentration for producing a specified effect in 50 percent of the
test organisms
EEC: Estimated Environmental Concentration
EPA: Environmental Protection Agency
FS: Feasibility Study
FSP: Field Sampling Plan
FWS: Fish and Wildlife Service
HEAST: National Center for Environmental Assessment's Health Effects Assessment
Summary Tables
HI: Hazard Index
HQ: Hazard Quotient
HSDB: National Library of Medicine's Hazardous Substances Data Bank
IRIS: EPA's Integrated Risk Information System
LC50: Concentration Lethal to 50 percent of the test organisms
Li " Liter
LOAEL: Lowest-Observed-Adverse-Effect Level
NCP: National Oil and Hazardous Substances Pollution Contingency Plan
NOAA: National Oceanic and Atmospheric Administration
NOAEL: No-Observed-Adverse-Effect Level
NRC: National Research Council
NRDA: Natural Resource Damage Assessment
OERR: U.S. EPA Office of Emergency and Remedial Response
OSC: On-Scene Coordinator
OSWER: U.S. EPA Office of Solid Waste and Emergency Response
PA Preliminary Assessment
PAH: Polycyclic Aromatic Hydrocarbons
PC^: Polychlorinated Biphenyl compound
PRP: Potentially Responsible Party
QAPP: Quality Assurance Project Plan
QA/QC: Quality Assurance and Quality Control
XIII
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RBP: Rapid Bioassessment Protocol
RI: Remedial Investigation
ROD: Record of Decision
RPM: Remedial Project Manager
SAP: Sampling and Analysis Plan
SARA: Superfund Amendments and Reauthorization Act of 1986
SI: Site Investigation
SMDP: Scientific/Management Decision Point
TOC: Total Organic Carbon
WP: Work Plan
XIV
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PREFACE
This document provides guidance on the process of designing and conducting
technically defensible ecological risk assessments for the Superfund Program. It is intended
to promote consistency and a science-based approach within the Program and is based on the
Proposed Guidelines for Ecological Risk Assessment (1996a) and the Framework for
Ecological Risk Assessment (1992a) developed by the Risk Assessment Forum of the U.S.
Environmental Protection Agency. When the Agency publishes its final Guidelines for
Ecological Risk Assessment, ibis guidance will be reviewed and revised if necessary to ensure
consistency with the Agency guidelines.
This document is directed to the site managers (i.e., On-Scene Coordinators [OSCs]
and Remedial Project Managers [RPMs]) who are legally responsible for the management of a
site. However, it is anticipated that ecological risk assessors, as well as other individuals with
input to the ecological risk assessment, will use this document. -
Ecological risk assessment is an integral pan of the Remedial Investigation and
Feasibility Study (RI/FS) process, which is designed to support risk management decision-
making for Superfund sites. The RI component of the process characterizes the nature and
extent of contamination at a hazardous waste site and estimates risks to human health and the
environment posed by contaminants at the site. The FS component of the process develops
and evaluates remedial options. Thus, ecological risk assessment is fundamental to the RI
and ecological considerations are also part of the FS process.
This document is intended to facilitate defensible site-specific ecological risk
assessments. It is not intended to determine the appropriate scale or complexity of an
ecological risk assessment or to direct the user in the selection of specific protocols or
investigation methods. Professional judgment is essential in designing and determining the
data needs for any ecological risk assessment. However, when the process outlined in this
document is followed, a technically defensible and appropriately scaled site-specific
ecological risk assessment should result.
Ecological risk assessment is an interdisciplinary field drawing upon environmental
toxicology, ecology, and environmental chemistry, as well as other areas of science and
mathematics. It is important that users of this document understand that ecological risk
assessment is a complex, non-linear process, with many parallel activities. The user should
have a basic understanding of ecotoxicology and ecological risk assessment and read through
this document in its entirety prior to engaging in the ecological risk assessment process.
Without the basic understanding of the field and of this guidance, the reader might not
recognize the relationships among different components of the risk assessment process.
To assist the user in interpreting this guidance document, three illustrations of
planning an ecological risk assessment for a hazardous waste site are provided in
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Appendix A. These are simplified, hypothetical examples that demonstrate and highlight
specific points in the ecological risk assessment process. These examples are incomplete and
not intended to present a thorough discussion of the ecological or ecotoxicological issues that
would exist at an actual site. Instead, they are intended to illustrate the first five steps of the
process, which precede a full ecological field investigation. Excerpts from the three examples
are included in the guidance document as "Example" boxes to illustrate specific points. The
user is encouraged to read the three examples in Appendix A in addition to the Example
boxes within the guidance document itself.
Ecological risk assessment is a dynamic field, and this document represents a process
framework into which changes in ecological risk assessment approaches can readily be
incorporated. Four appendices are included with this document; additional appendices may be
developed to address specific issues.
This document supersedes the U.S. EPA's (1989b) Risk Assessment Guidance for
Superfund, Volume 2: Environmental Evaluation Manual as guidance on how to design and
conduct an ecological risk assessment for the Superfund Program. The Environmental
Evaluation Manual contains useful information on the statutory and regulatory basis of
ecological assessment, basic ecological concepts, and other background information that is not
repeated in this document.
XVI
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INTRODUCTION:
ECOLOGICAL RISK ASSESSMENT FOR SUPERFUND
PURPOSE
This document provides guidance on how to design and conduct consistent and
technically defensible ecological risk assessments for the Superfund Program. It is based on
the Proposed Guidelines for Ecological Risk Assessment (1996a) and the Framework for
Ecological Risk Assessment (1992a) developed by the Risk Assessment Forum of the U.S.
Environmental Protection Agency (U.S. EPA or the Agency). When the Agency finalizes its
(1996a) Proposed Guidelines for Ecological Risk Assessment, this guidance will be reviewed
and revised if necessary to ensure consistency with the Agency guidelines.
This document is directed to the site managers (i.e., On-Scene Coordinators [OSCs]
and Remedial Project Managers [RPMsJ) who are legally responsible for managing site
activities. However, it is anticipated that the ecological risk assessors, as well as all other
individuals involved with ecological risk assessments, will use this document.
SCOPE
This document is intended to facilitate defensible and appropriately-scaled site-specific
ecological risk assessments. It is not intended to dictate the scale, complexity, protocols, data
needs, or investigation methods for such assessments. Professional judgment is required to
apply the process outlined in this document to ecological risk assessments at specific sites.
BACKGROUND
Superfund Program
The Comprehensive Environmental Response, Compensation, and Liability Act of
1980 (CERCLA or Superfund), as amended by the Superfund Amendments and
Reauthorization Act of 1986 (SARA), authorizes the U.S. EPA to protect public health and
welfare and the environment from the release or potential release of any hazardous substance,
pollutant, or contaminant. U.S. EPA's Superfund Program carries out the Agency's mandate
under CERCLA/SARA.
The primary regulation issued by U.S. EPA's Superfund Program is the National Oil
and Hazardous Substances Pollution Contingency Plan (NCP). The NCP calls for the
identification and mitigation of environmental impacts (such as toxicity, bioaccumulation,
death, reproductive impairment, growth impairment, and loss of critical habitat) at hazardous
waste sites, and for the selection of remedial actions to protect the environment. In addition,
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numerous other federal and state laws and regulations concerning environmental protection
can be designated under Superfund as "applicable" or "relevant and appropriate" requirements
(ARARs) for particular sites. Compliance with these other laws and regulations generally
requires an evaluation of site-related ecological effects and the measures necessary to mitigate
those effects.
Risk Assessment in Superfund
An important part of the NCP is the
requirement for a Remedial Investigation
and Feasibility Study (RI/FS) (see Highlight
1-1). The RI^S is an analytical process
designed to support risk management
decision-making for Superfund sites. The
RI component of the process characterizes
the nature and extent of contamination at a
hazardous waste site and estimates risks to
human health and the environment posed by
contaminants at the site. The FS component
of the process develops and evaluates
remedial options.
HIGHLIGHT 1-1
The RI/FS Process
Risk assessment is an integral part of
the RI/FS. The three pans of the RI are: (!)
characterization of the nature and extent of
contamination; (2) ecological risk
assessment; and (3) human health risk
assessment. The investigation of the nature
and extent of contamination determines the
chemicals present on site as well as their
distribution and concentrations. The
ecological risk and human health risk
assessments determine the potential for
adverse effects to the environment and
human health, respectively.
Although U.S. EPA has established
detailed guidelines for human health risk
assessment in the Superfund program (U.S.
EPA, 1989a, 1991a,b), similarly detailed guidelines for site-specific ecological risk assessment
do not exist for the Superfund program. Risk Assessment Guidance for Superfund, Volume 2:
Environmental Evaluation Manual (U.S. EPA, 1989b) provides conceptual guidance in
planning studies to evaluate a hazardous waste site's "environmental resources" (as used in
the manual, the phrase "environmental resources" is largely synonymous with "ecological
resources"). U.S. EPA also is publishing supplemental information on specific ecological risk
assessment topics for Superfund in the ECO Update series (U.S. EPA, 1995b, I994b,c,d,e,
1992b,c,d, 1991c,d). However, those documents do not describe an overall, step-by-step
process by which an ecological risk assessment is designed and executed. The Agency's
Framework for Ecological Risk Assessment (U.S. EPA, 1992a) provides a basic structure and
a consistent approach for conducting ecological risk assessments, but is not intended to
provide program-specific guidance. The Guidelines for Ecological Risk Assessment, currently
being developed by the Agency's Risk Assessment Forum (1996a), will expand on the
Framework, but again, will not provide program-specific guidance.
This document outlines a step-by-step ecological risk assessment process that is both
specific to the Superfund Program and consistent with the more general U.S. EPA Framework
and guidelines under development. While the Agency's Framework and future Agency-wide
ecological risk assessment guidelines are not enforceable regulations, the concepts in those
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documents are appropriate to Superfund. The concepts in the published Framework have
been incorporated into this document with minimal modification. The definitions of terms
used in this ecological risk assessment guidance for Superfund (and listed in the Glossary) are
consistent with the definitions in the U.S. EPA Framework document unless noted otherwise.
DEFINITION OF ECOLOGICAL RISK ASSESSMENT
U.S. EPA "Framework" Document
Ecological risk assessment is defined in the Framework as a process that evaluates the
likelihood that adverse ecological effects are occurring or may occur as a result of exposure
to one or more stressors (U.S. EPA, 1992a). The Framework defines a stressor as any
physical, chemical, or biological entity that can induce an adverse ecological response.
Adverse responses can range from Sublethal chronic effects in individual organisms to a loss
of ecosystem function. Although stressors can be biological (e.g., introduced species), only
chemical or physical stressors will be addressed in this document, because these are the
stressors subject to risk management decisions at Superfund sites.
Superfund Program
The phrase "ecological risk assessment," as used specifically for the Superfund
Program in this document, refers to a qualitative and/or quantitative appraisal of the actual or
potential impacts of contaminants from a hazardous waste site on plants and animals other
than humans and domesticated species. A risk does not exist unless: (1) the stressor has the
ability to cause one or more adverse effects, and (2) it co-occurs with or contacts an
ecological component long enough and at a sufficient intensity to elicit the identified adverse
effect.
THE ECOLOGICAL RISK ASSESSMENT PROCESS
U.S. EPA "Framework" Document
The Framework describes the basic elements of a process for scientifically evaluating
the adverse effects of stressors on ecosystems and components of ecosystems. The document
describes the basic process and principles to be used in ecological risk assessments conducted
for the U.S. EPA, provides operational definitions for terms used in ecological risk
assessments, and outlines basic principles around which program-specific guidelines for
ecological risk assessment should be organized.
The Framework is similar to the National Research Council's (NRC) paradigm for
human health risk assessments (NRC, 1983) and the more recent NRC ecological risk
paradigm (NRC, 1993). The 1983 NRC paradigm consists of four fundamental phases:
1-3
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hazard identification, dose-response assessment, exposure assessment, and risk
characterization. The Framework differs from the 1983 NRC paradigm in a few ways:
• Problem formulation is incorporated into the beginning of the process to .
determine the focus and scope of the assessment;
• Hazard identification and dose-response assessment are combined in an .
ecological effects assessment phase; and
• The phrase "dose-response" is replaced by "stressor-response" to emphasize the
possibility that physical changes (which are not measured in "doses") as well as
chemical contamination can stress ecosystems.
Moreover, the Framework emphasizes the parallel nature of the ecological effects and
exposure assessments by joining the two assessments in an .analysis phase between problem
formulation and risk characterization, as shown in Exhibit 1-1.
During problem formulation, the risk assessor establishes the goals, breadth, and focus
of the assessment (U.S. EPA, 1992a). As indicated in the Framework, problem formulation is
a systematic planning step that identifies the major factors to be considered and is linked to
the regulatory and policy contexts of the assessment. Problem formulation includes .
discussions between the risk assessor and risk manager, and other involved parties, to identify
the stressor characteristics, ecosystems potentially at risk, and ecological effects to be
evaluated. During problem formulation, assessment and measurement endpoints for the
ecological risk assessment are identified, as described below.
The Agency defines assessment endpoints as explicit expressions of the actual
environmental values (e.g., ecological resources) that are to be protected (U.S. EPA, 1992a).
Valuable ecological resources include those without which ecosystem function would be
significantly impaired, those providing critical resources (e.g., habitat, fisheries), and those
perceived as valuable by humans (e.g., endangered species and other issues addressed by
legislation). Because assessment endpoints focus the risk assessment design and analysis,
appropriate selection and definition of these endpoints are critical to the utility of a risk
assessment.
Assessment endpoints should relate to statutory mandates (e.g., protection of the
environment), but must be specific enough to guide the development of the risk assessment
study design at a particular site. Useful assessment endpoints define both the valued
ecological entity at the site (e.g., a species, ecological resource, or habitat type) and a
characteristic(s) of the entity to protect (e.g., reproductive success, production per unit area,
areal extent). Highlight 1-2 provides some examples of specific assessment endpoints related
to the general goal of protecting aquatic ecosystems.
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EXHIBIT 1-1
Ecological Risk Assessment Framework (U.S. EPA, 1992a)
Discussion
Between the
Risk Assessor
and
Risk Manager
(Planning)
Ecological Risk Assessment
PROBLEM FORMULATION
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1-5
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biological response to a stressor that can be
assessment endpoint (U.S. EPA, 1992a;
A measurement endpoint is a measurable
related to the valued characteristic chosen as the
although this definition may change—see
U.S. EPA, 1996a). Sometimes, the
assessment endpoint can be measured
directly; usually, however, an assessment
endpoint encompasses too many species or
species that are difficult to evaluate (e.g.,
top-level predators). In these cases, the
measurement endpoints are different from
the assessment endpoint, but can be used to
make inferences about risks to the
assessment endpoints. For example,
measures of responses in particularly
sensitive species and life stages might be
used to infer responses in the remaining
species and life stages in a specific
community. Such inferences must be
clearly described to demonstrate the link
between measurement and assessment
endpoints. Highlight 1-3 provides examples
of measurement endpoints.
Measures of exposure also can be used to make inferences about risks to assessment
endpoints at Superfund sites. For example, measures of water concentrations of a
contaminant can be compared with concentrations known from the literature to be lethal to
sensitive aquatic organisms to infer something about risks to aquatic community structure. As
a consequence, for purposes of this guidance, measurement endpoints include both measures
of effect and measures of exposure.
HIGHLIGHT 1-2
Example Assessment Endpoints
• Sustained aquatic community
structure, including species
composition and relative abundance
and trophic structure.
• Sufficient rates of survival, growth,
and reproduction to sustain
populations of carnivores typical for
the area.
• Sustained fishery diversity and
abundance.
A product of problem formulation is
a conceptual model for the ecological risk
assessment that describes how a given
stressor might affect ecological components
of the environment. The conceptual model
also describes questions about how stressors
affect the assessment endpoints, the
relationships among the assessment and
measurement endpoints, the data required to
answer the questions, and the methods that
will be used to analyze the data (U.S. EPA,
1992a).
HIGHLIGHT 1-3
Example Measurement Endpoints
• Community analysis of benthic
macroin vertebrates.
• Survival and growth of fish fry in
response to exposure to copper.
• Community structure of fishery in
proximity to the site.
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Superfund Program
The goal of the ecological risk assessment process in the Superfund Program is to
provide the risk information necessary to assist risk managers at Superfund sites (OSCs and
RPMs) in making informed decisions regarding substances designated as hazardous under
CERCLA (see 40 CFR 302.4). The specific objectives of the process, as stated in OSWER
Directive 9285.7-17, are: (1) to identify and characterize the current and potential threats to
the environment from a hazardous substance release; and (2) to identify cleanup levels that
would protect those natural resources from risk. Threats to the environment include existing
adverse ecological impacts and the risk of such impacts in the future. Highlight 1-4 provides
an overview of ecological risk assessment in the Superfund Program.
Problem formulation is the most critical step of an ecological risk assessment and must
precede any attempt to design a site investigation and analysis plan. To ensure that the risk
manager can use the results of an ecological risk assessment to inform risk management
decisions for a Superfund site, it is important that all involved parties contribute to the
problem formulation phase and that the risk manager is clearly identified to all parties. These
parties include the remedial project manager (RPM), who is the risk manager with ultimate
responsibility for the site, the ecological risk assessment team, the Regional Superfund
Biological Technical Assistance Group (STAG), potentially responsible parties (PRPs),
Natural Resource Trustees, and stakeholders in the natural resources at issue (e.g., local
communities, state agencies) (U.S. EPA, 1994a, 1995b). The U.S. EPA's (1994a) Edgewater
Consensus on an EPA Strategy for Ecosystem Protection in particular calls for the Agency to
develop a "place-driven" orientation, that is, to focus on the environmental needs of specific
communities and ecosystems, rather than on piecemeal program mandates. Participation in
problem formulation by all involved parties helps to achieve the place-driven focus.
Issues such as restoration, mitigation, and replacement are important to the Superfund
Program; but are reserved for investigations that might or might not be included in the RI
phase. During the risk management process of selecting the preferred remedial option leading
to the Record of Decision (ROD), issues of mitigation and restoration should be addressed.
In selecting a remedy, the risk manager must also consider the degree to which the remedial
alternatives reduce risk and thereby also reduce the need for restoration or mitigation.
A natural resource damage assessment (NRDA) may be conducted at a Superfund site
at the discretion of Natural Resource Trustees for specific resources associated with a site.
An ecological risk assessment is a necessary step for an NRDA, because it establishes the
causal link between site contaminants and specific adverse ecological effects. The risk
assessment also can provide information on what residual risks are likely for different
remediation options. However, the ecological risk assessment does not constitute an NRDA.
The NRDA is the sole responsibility of the Natural Resource Trustees, not of the U.S. EPA;
therefore, NRDAs will not be addressed in this guidance. For additional information on the
role of Natural Resource Trustees in the Superfund process, see ECO Update Volume 1,
Number 3 (U.S. EPA, 1992c).
1-7
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HIGHLIGHT 1-4
Ecological Impact and Risk Assessment
Ecological risk assessment within the Superfund Program can be a risk evaluation
(potentially predictive), impact evaluation, or a combination of those approaches. The
functions of the ecological risk assessment are to:
(1) Document whether actual or potential ecological risks exist at a site;
(2) Identify which contaminants present at a site pose an ecological risk; and
(3) Generate data to be used in evaluating cleanup options.
Ecological risk assessments can have their greatest influence on risk management at a site in
the evaluation and selection of site remedies. The ecological risk assessment should identify
contamination levels that bound a threshold for adverse effects on the assessment endpoint.
The threshold values provide a yardstick for evaluating the effectiveness of remedial options
and can be used to set cleanup goals if appropriate.
To justify a site action based upon ecological concerns, the ecological risk assessment
must establish that an actual or potential ecological threat exists at a site. The potential for
(i.e., risk of) impacts can be the threat of impacts from a future release or redistribution of
contaminants, which could be avoided by taking actions on "hot spots" or source areas. Risk
also can be viewed as the likelihood that current impacts are occurring (e.g., diminished
population size), although this can be difficult to demonstrate. For example, it may not be
practical or technically possible to document existing ecological impacts, either due to limited
technique resolution, the localized nature of the actual impact, or limitations resulting from
the biological or ecological constraints of the field measurements (e.g., measurement
endpoints, exposure point evaluation). Actually demonstrating existing impacts confirms that
a "risk" exists. Evaluating a gradient of existing impacts along a gradient of contamination
can provide an stressor-response assessment that helps to identify cleanup levels.
As noted above, the ecological risk assessment should provide the information needed
to make risk management decisions (e.g., to select the appropriate site remedy). A
management option should not be selected first, and then the risk assessment tailored to
justify the option.
This Guidance Document
This ecological risk assessment guidance for Superfund is composed of eight steps
(see Exhibit 1-2) and several scientific/management decision points (SMDPs) (see Exhibit
1-3). An SMDP requires a meeting between the risk manager and risk assessment team to
evaluate and approve or redirect the work up to that point. (Consultation with the Regional
BTAG is recommended for SMDPs (a) through (d) in Exhibit 1-3.) The group decides
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EXHIBIT 1-2
Eight-step Ecological Risk Assessment Process for Superfund
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STEP1: SCREENING-LEVEL
• Site Visit
• Problem Formulation
• Toxicity Evaluation
STEP 2: SCREENING-LEVEL:
• Exposure Estimate
• Risk Calculation
STEP 3: PROBLEM FORMULATION
Toxicity Evaluation
Assessment
Endpoints
I
Conceptual Model
Exposure Pathways
Questions/Hypotheses
STEP 5: VERIFICATION OF FIELD
SAMPLING DESIGN
STEP 6: SITE INVESTIGATION AND
DATA ANALYSIS
STEP 7: RISK CHARACTERIZATION
Risk Assessor
and Risk Manager
Agreement
i
SMDP
STEP 4: STUDY DESIGN AND DQO PROCESS
• Lines of Evidence
• Measurement Endpoints
Work Plan and Sampling and Analysis Plan
SMDP
STEP 8:
RISK MANAGEMENT
SMDP ,
1-9
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EXHIBIT 1-3
Steps in the Ecological Risk Assessment Process
and Corresponding Decision Points in the Superfund Process
Steps and Scientific/Management Decision Points (SMDPs):
1. Screening-Level Problem Formulation and Ecological
Effects Evaluation
2. Screening-Level Preliminary Exposure Estimate and
Risk Calculation SMDP (a)
3. Baseline Risk Assessment Problem Formulation SMDP (b)
4. Study Design and Data Quality Objectives SMDP (c)
5. Field Verification of Sampling Design SMDP (d)
6. Site Investigation and Analysis of Exposure
and Effects , [SMDP]
7. Risk Characterization
8. Risk Management SMDP (e)
Corresponding Decision Points in the Superfund Process:
(a) Decision about whether a full ecological risk assessment
is necessary.
(b) Agreement among the risk assessors, risk manager, and
other involved panics on the conceptual model,
including assessment endpoints, exposure pathways, and
.questions or risk hypotheses.
(c) Agreement among the risk assessors and risk manager on the
measurement endpoints, study design, and data interpretation
and analysis.
(d) Signing approval of the work plan and sampling and analysis
plan for the ecological risk assessment.
(e) Signing the Record of Decision.
[SMDP] only if change to the sampling and analysis plan is necessary.
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whether or not the risk assessment is proceeding in a direction that is acceptable to the risk
assessors and manager. The SMDPs include a discussion of the uncertainty associated with
the risk assessment, that might be reduced, if necessary, with increased effort. SMDPs are
significant communication points which should be passed with the consensus of all involved
parties. The risk manager should expect deliverables that document specific SMDPs as
outlined in Exhibit 1-4. This approach is intended to minimize both the cost of and time
required for the Superfund risk assessment process.
This guidance provides a technically valid approach for ecological risk assessments at
hazardous waste sites, although other approaches also can be valid. The discipline of
ecological risk assessment is dynamic and continually evolving; the assessments rely on data
that are complex and sometimes ambiguous. Thus, if an approach other than the one
described in this guidance document is used, there must be clear documentation of the
process, including process design and interpretation of the results, to ensure a technically
defensible assessment. Clear documentation, consistency, and objectivity in the assessment
process are necessary for the Superfund Program.
An interdisciplinary team including, but not limited to, biologists, ecologists, and
environmental lexicologists, is needed to design and implement a successful risk assessment
and to evaluate the weight of the evidence obtained to reach conclusions about ecological
risks. Some of the many points at which the Superfund ecological risk assessment process
requires professional judgment include:
EXHIBIT 1-4
Ecological Risk Assessment Deliverables
for the Risk Manager
If the process stops at the end of Step 2:
(1) Full documentation of the screening-level assessment and SMDP not to continue
the assessment.
If the process continues to Step 3:
(1) Documentation of the conceptual model, including assessment endpoints,
exposure pathways, risk hypotheses, and SMDP at the end of Step 3.
(2) The approved and signed work plan and sampling and analysis plan,
documenting the SMDPs at the end of Steps 4 and 5.
(3) The baseline risk assessment documentation (including documentation of the
screening-level assessment used in the baseline assessment) developed in Step 7.
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• Determining the level of effort needed to assess ecological risk at a particular
site;
• Determining the relevance of available data to the risk assessment;
• Designing a conceptual model of the ecological threats at a site and measures
to assess those threats;
• Selecting methods and .models to be used in the various components of the risk
assessment;
• Developing assumptions to fill data gaps for toxicity and exposure assessments
based on logic and scientific principles; and
• Interpreting the ecological significance of observed or predicted effects.
The lead risk assessor should coordinate with appropriate professionals to make many of these
decisions.. Specialists are needed for the more technical questions concerning the risk
assessment (e.g., which model, which assumptions).
This guidance document focuses on the risk assessment process in Superfund and does
not address all of the issues that a risk manager will need to consider. After the risk
assessment is complete, the risk manager might require additional professional assistance in
interpreting the implications of the baseline ecological risk assessment and selecting a
remedial option.
The risk assessment process must be structured to ensure that site management
decisions can be made without the need for repeated studies or delays. The first two steps in
the assessment process are a streamlined version of the complete Framework process and are
intended to allow a rapid determination by the risk assessment team and risk manager that the
site poses no or negligible 'ecological risk, or to identify which contaminants and exposure
pathways require further evaluation. Steps 3 through 7 are a more detailed version of the
complete Framework process.
The ecological risk assessment process should be coordinated with the overall RI/FS
process to the extent possible. Overall site-assessment costs are minimized when the needs of
the ecological and human health risk assessments are incorporated into the chemical sampling
program to determine the nature and extent of contamination during the RI. For sites at
which an RI has not yet been planned or conducted, Exhibit 1-5 illustrates the relationship
between the eight ecological risk assessment steps and the overall Superfund process and
decision points. For older sites at which an RI was conducted before an ecological risk
assessment was considered, the ecological risk assessment process should build on the
information already developed for the site.
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EXHIBIT 1-5
Ecological Risk Assessment in the RI/FS Process
FROM:
• Preliminary Assessment
• Site Inspection
• NPL Listing
Remedial Investigation
WP
and
SAP
Site
Investigation
Feasibility Study
Establish
Remedial
Objectives
SREENINQ
ECOLOGICAL RISK
ASSESSMENT
(STEPS 1 & 2)
FIELD
VERIFICATION
(STEP 5)
PROBLEM
FORMULATION AND
STUDY DESIGN
(STEPS 3 & 4)
Refine remedial
goals based on
risk assessment
ANALYSIS OF
EXPOSURE AND EFFECTS
RISK CHARACTERIZATION
(STEPS 6 » 7)
Development
and Analysis
of Alternatives
TO:
• Remedy Selection
• Record of Decision
• Remedial Design
• Remedial Action
Conduct risk
evaluation of
remedial
alternatives
Ecological
Monitoring
-------
It is important to realize that this eight-step approach is not a simple linear or
sequential process. The order of actions taken will depend upon the stage of the RI/FS
atwhich the site is currently, the amount and types of site information available, as well as
other factors. The process can be iterative, and in some iterations, certain individual steps
might not be needed. In many cases, it might be appropriate and desirable to conduct several
steps concurrently.
Tasks that should be accomplished in each of the eight steps in Exhibits 1-2 and 1-3
are described in the eight following sections. The eight sections include example boxes based
on the three hypothetical Superfund sites in Appendix A as well as exhibits and highlight
boxes.
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STEP 1: SCREENING-LEVEL PROBLEM FORMULATION
AND ECOLOGICAL EFFECTS EVALUATION
OVERVIEW
The screening-level problem formulation and ecological effects evaluation is
pan of the initial ecological risk screening assessment. For this initial step, it is likely
that site-specific information for determining the nature and extent of contamination
and for characterizing ecological receptors at the site is limited. This step includes all
the functions of problem formulation (more fully described in Steps 3 and 4) and
ecological effects analysis, but on a screening level. The results of this step will be
used in conjunction with exposure estimates in the preliminary risk calculation in
Step 2.
1.1 INTRODUCTION
Step 1 is the screening-level problem formulation process and ecological effects
evaluation (Highlight 1-1 defines screening-level risk assessments). Consultation with the
STAG is recommended at this stage. How to brief the BTAG on the setting, history, and
ecology of a site is described in ECO Update Volume 1, Number 5 (U.S. EPA, 1992d).
Section 1.2 describes the screening-level problem formulation, and Section 1.3 describes the
screening-level ecological effects evaluation. Section 1.4 summarizes this step.
1.2 SCREENING-LEVEL PROBLEM FORMULATION
/
For the screening-level problem formulation, the risk assessor develops a conceptual
model for the site that addresses five issues:
(1) Environmental setting and contaminants known or suspected to exist at the site
(Section 1.2.1);
(2) Contaminant fate and transport mechanisms that might exist at the site (Section
1.2.2);
(3) The mechanisms of ecotoxicity associated with contaminants and likely categories
of receptors that could be affected (Section 1.2.3);
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(4) What complete exposure
pathways might exist at the site
(a complete exposure pathway is
one in which the chemical can
be traced or expected to travel
from the source to a receptor
that can be affected by the
chemical) (Section 1.2.4); and
(5) Selection of endpoints to screen
for ecological risk (Section
1.2.5).
1.2.1 Environmental Setting and
Contaminants at the Site
HIGHLIGHT 1-1
Screening-level Risk Assessments
Screening-level risk assessments are
simplified risk assessments -that can be
conducted with limited data by assuming
values for parameters for which data are
lacking. At the screening level, it is
important to minimize the chances of
concluding that there is no risk when in fact
a risk exists. Thus, for exposure and toxicity
parameters for which site-specific information
is lacking, assumed values should
consistently be biased in the direction of
overestimating risk. This ensures that sites
that might pose an ecological risk are studied
further. Without this bias, a screening
evaluation could not provide a defensible
conclusion that negligible ecological risk
exists or that certain contaminants and
exposure pathways can be eliminated from
consideration.
To begin the screening-level
problem formulation, there must be at least
a rudimentary knowledge of the potential
environmental setting and chemical
contamination at the site. The first step is
to compile information from the site history
and from reports related to the site,
including the Preliminary Assessment (PA)
or Site Investigation (SI). The second step is to use the environmental checklist presented in
Representative Sampling Guidance Document. Volume 3: Ecological (U.S. EPA, 1997; see
Appendix B) to begin characterizing the site for problem, formulation. Key questions
addressed by the checklist include:
• What are the on- and off-site land uses (e.g., industrial, residential, or
undeveloped; current and future)?
• What type of facility existed or exists at the site?
• What are the suspected contaminants at the site?
• What is the environmental setting, including natural areas (e.g., upland forest,
on-site stream, nearby wildlife refuge) as well as disturbed/man-made areas
(e.g., waste lagoons)?
• Which habitats present on site are potentially contaminated or otherwise
disturbed?
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• Has contamination migrated from source areas and resulted in "off-site"
impacts or the threat of impacts in addition to on-site threats or impacts?
These questions should be answered using the site reports, maps (e.g. U.S. Geological
Survey, National Wetlands Inventory), available aerial photographs, communication with
appropriate agencies (e.g., U.S. Fish and Wildlife Service, National Oceanic and Atmospheric
Administration, State Natural Heritage Programs), and a site visit. Activities that should be
conducted during the site visit include:
• Note the layout and topography of the site;
• Note and describe any water bodies and wetlands;
• Identify and map evidence indicating contamination or potential contamination
(e.g., areas of no vegetation, runoff gullies to surface waters);
• Describe existing aquatic, terrestrial, and wetland ecological habitat types (e.g.,
forest, old field), and estimate the area covered by those habitats;
• Note any potentially sensitive environments (see Section 1.2.3 for examples of
sensitive environments);
• Describe and, if possible, map soil and water types, land uses, and the
dominant vegetation species present; and
• Record any observations of animal species or sign of a species.
Mapping can be useful in establishing a "picture" of the site to assist in problem
formulation. The completed checklist (U.S. EPA, 1997) will provide information regarding
habitats and species potentially or actually present on site, potential contaminant migration
pathways, exposure pathways, and the potential for non-chemical stresses at the site.
After finishing the checklist, it might be possible to determine that present or future
ecological impacts are negligible because complete exposure pathways do not exist and could
not exist in the future. Many Superfund sites are located in highly industrialized areas where
there could be few if any ecological receptors or where site-related impacts might be
indistinguishable from non-site-related impacts (see Highlight 1-2). For such sites,
remediation to reduce ecological risks might not be needed. However, all sites should be
evaluated by qualified personnel to determine whether this conclusion is appropriate:
Other Superfund sites are located in less disturbed areas with protected or sensitive
environments that could be at risk of adverse effects from contaminants from the site. State
and federal laws (e:g., the Clean Water Act, the Endangered Species Act) designate certain
types of environments as requiring protection. Other types of habitats unique to certain areas
1-3
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HIGHLIGHT 1-2
Industrial or Urban Settings
Many hazardous waste sites exist
in currently or historically industrialized
or urbanized areas. In these instances, it
can be difficult to distinguish between
impacts related to contaminants from a
particular site and impacts related to
non-contaminant stressors or to
contaminants from other sites. However,
even in these cases, it could be
appropriate to take some remedial
actions based on ecological risks. These
actions might be limited to source
removal or might be more extensive.
An ecological risk assessment can assist
the risk manager in determining what
action, if any, is appropriate.
also could need special consideration in the risk
assessment (see Section 1.2.3).
1.2.2 Contaminant Fate and Transport
During problem formulation, pathways
for migration of a contaminant (e.g., windblown
dust, surface water runoff, erosion) should be
identified. These pathways can exhibit a
decreasing gradient of contamination with
increasing distance from a site. There are
exceptions, however, because physical and
chemical characteristics of the media also
influence contaminant distribution (e.g., the
pattern of sediment deposition in streams varies
depending on stream flow and bottom
characteristics). For the screening-level risk
assessment, the highest contaminant
concentrations measured on the site should be
documented for each medium.
1.2.3 Ecotoxicity and Potential Receptors
Understanding the toxic mechanism of a contaminant helps to evaluate the importance
of potential exposure pathways (see Section 1.2.4) and to focus the selection of assessment
endpoints (see Section 1.2.5). Some contaminants, for example, affect primarily vertebrate
animals by interfering with organ systems not found in invertebrates or plants (e.g., distal
tubules of vertebrate kidneys, vertebrate hormone systems). Other substances might affect
primarily certain insect groups (e.g., by interfering with hormones needed for metamorphosis),
plants (e.g., herbicides), or other groups of organisms. For substances that affect, for
example, reproduction of mammals at much lower environmental exposure levels than they
affect other groups of organisms, the screening-level risk assessment can initially focus on
exposure pathways and risks to mammals. Example 1-1 illustrates this point using the PCB
site example provided in Appendix A. A review of some of the more recent ecological risk
and toxicity assessment literature can help identify likely effects of the more common
contaminants at Superfund sites.
An experienced biologist or ecologist can determine what plants, animals, and habitats
exist or can be expected to exist in the area of the Superfund site. Exhibit 1-1, adapted from
the Superfund Hazard Ranking System, is a partial list of types of sensitive environments that
could require protection or special consideration. Information obtained for the environmental
checklist (Section 1.2.1), existing information and maps, and aerial photographs should be
used to identify the presence of sensitive environments on or near a site that might be
threatened by contaminants from the site.
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EXAMPLE 1-1
Ecotoxicity-PCB Site
Some PCBs are reproductive toxins in mammals (Ringer et al., 1972; Aulerich et al.,
1985; Wren et al., 1991; Kamrin and Ringer, 1996). When ingested, they induce (i.e., increase
concentrations and activity of) enzymes in the liver, which might affect the metabolism of some
steroid hormones (Rice and O'Keefe, 1995). Whatever the mechanism of action, several
physiological functions that are controlled by steroid hormones can be altered by the exposure
of mammals to certain PCBs, and reproduction appears to be the most sensitive endpoint for
PCB toxicity in mammals (Rice and O'Keefe, 1995). Given this information, the screening
ecological risk assessment should include potential exposure pathways for mammals to PCBs
that are reproductive toxins (see Example 1-2).
1.2.4 Complete Exposure Pathways
Evaluating potential exposure pathways is one of the primary tasks of the screening-
level ecological characterization of the site. For an exposure pathway to be complete, a
contaminant must be able to travel from the source to ecological receptors and to be taken up
by the receptors via one or more exposure routes. (Highlight 1-3 defines exposure pathway
and exposure route.) Identifying complete exposure pathways prior to a quantitative
evaluation of toxicity allows the assessment to focus on only those contaminants that can
reach ecological receptors.
Different exposure routes are important for different groups of organisms. For
terrestrial animals, three basic exposure routes need to be evaluated: inhalation, ingestion,
and dermal absorption. For terrestrial plants, root absorption of contaminants in soils and leaf
absorption of contaminants evaporating from the soil or deposited on the leaves are of
concern at Superfund sites. For aquatic animals, direct contact (of water or sediment with the
gills or integument) and ingestion of food (and sometimes sediments) should be considered.
For aquatic plants, direct contact with water, and sometimes with air or sediments, is of
primary concern.
The most likely exposure pathways and exposure routes also are related to the physical
and chemical properties of the contaminant (e.g., whether or not the contaminant is bound to
a matrix, such as organic carbon). Of the basic exposure routes identified above, more
information generally is available to quantify.exposure levels for ingestion by terrestrial
animals and for direct contact with water or sediments by aquatic organisms than for other
exposure routes and receptors. Although other exposure routes can be important, more
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EXHIBIT 1-1
List of Sensitive Environments in the Hazard Ranking System8
Critical habitat for Federal designated endangered or threatened species
Marine Sanctuary
National Park
Designated Federal Wilderness Area
Areas identified under the Coastal Zone Management Act . •
Sensitive areas identified under the National Estuary Program or Near Coastal Waters Program
Critical areas identified under the Clean Lakes Program
National Monument
National Seashore Recreational Area
National Lakeshore Recreational Area
Habitat known to be used by Federal designated or proposed endangered or threatened species
National Preserve
National or State Wildlife Refuge
Unit of Coastal Barrier Resources System
Coastal Barrier (undeveloped)
Federal land designated for protection of natural ecosystems
Administratively Proposed Federal Wilderness Area
Spawning areas critical for the maintenance of fish/shellfish species within river, lake, or
coastal tidal waters
Migratory pathways and feeding areas critical for maintenance of anadromous fish species within river
reaches or areas in lakes or coastal tidal waters in which the fish spend extended periods of time
Terrestrial areas utilized for breeding by large or dense aggregations of animals
National river reach designated as Recreational
Habitat known to be used by state designated endangered or threatened species
Habitat known to be used by species under review as to its Federal endangered or threatened status
Coastal Barrier (partially developed)
Federally-designated Scenic or Wild River
State land designated for wildlife or game management
State-designated Scenic or Wild River -
State-designated Natural Areas
Particular areas, relatively small in size, important to maintenance of unique biotic communities
State-designated areas for protection or maintenance of aquatic life
Wetlands6
a The categories are listed in groups from those assigned higher factor values to those assigned
lower factor values in the Hazard Ranking System (HRS) for listing hazardous waste sites on the National
Priorities List (U.S. EPA, 1990b). See Federal Register, Vol. 55. pp. 51624 and 51648 for additional
information regarding definitions.
b Under the HRS. wetlands are rated on the basis of size. See Federal Register, Vol. 55, pp.
51625 and 51662 for additional information.
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HIGHLIGHT 1-3
Exposure Pathway and
Exposure Route
Exposure Pathway: The pathway by
which a contaminant travels from a source
(e.g., drums, contaminated soils) to
receptors. A pathway can involve multiple
media (e.g., soil runoff to surface waters and
sedimentation, or volatilization to the
atmosphere).
Exposure Route: A point of contact/entry
of a contaminant from the environment into
an organism (e.g., inhalation, ingestion,
dermal absorption).
assumptions are needed to estimate exposure
levels for those routes, and the results are
less certain. Professional judgment is
needed to determine if evaluating those
routes sufficiently improves a risk
assessment to warrant the effort.
If an exposure pathway is not
complete for a specific contaminant (i.e.,
ecological receptors cannot be exposed to
the contaminant), that exposure pathway
does not need to be evaluated further. For
example, suppose a contaminant that impairs
reproduction in mammals occurs only in
soils that are well below the root zone of
plants that occur or are expected to occur on
a site. Herbivorous mammals would not be
exposed to the contaminant through their
diets because plants would not be
contaminated. Assuming that most soil macroinvertebrates available for ingestion live in the
root zone, insectivorous mammals also would be unlikely to be exposed. In this case, a
complete exposure pathway for this contaminant for ground-dwelling mammals would not
exist, and the contaminant would not pose a significant risk to this group of organisms.
Secondary questions might include whether the contaminant is leaching from the soil to
ground water that discharges to surface water, thereby posing a risk to the aquatic
environment or to terrestrial mammals that drink the water or consume aquatic prey.
Example 1-2 illustrates the process of identifying complete exposure pathways based on the
hypothetical PCB site described in Appendix A.
1.2.5 Assessment and Measurement Endpoints
For the screening-level ecological risk assessment, assessment endpoints are any
adverse effects on ecological receptors, where receptors are plant and animal populations and
communities, habitats, and sensitive environments. Adverse effects on populations can be
inferred from measures related to impaired reproduction, growth, and survival. Adverse
effects on communities can be inferred from changes in community structure or function.
Adverse effects on habitats can be inferred from changes in composition and characteristics
that reduce the habitats1 ability to support plant and animal populations and communities.
Many of the screening ecotoxicity values now available or likely to be available in the
future for the Superfund program (see Section 1.3) are based on generic assessment endpoints
(e.g., protection of aquatic communities from changes in structure or function) and are
assumed to be widely applicable to sites around the United States.
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EXAMPLE 1-2
Complete Exposure Pathways for Mammals-PCB Site
Three possible exposure pathways for mammals were evaluated at the PCB Site:
inhalation, ingestion through the food chain, and incidental soil/sediment ingestion.
Inhalation. PCBs are not highly volatile, so the inhalation of PCB vapors by
mammals would be an essentially incomplete exposure pathway. Inhalation of PCBs adsorbed
to soil panicles might need consideration in areas with exposed soils, but this site is well
vegetated.
Ingestion through the food chain. PCBs tend to bioaccumulate and biomagnify in
food chains. PCBs in soils are not taken up by most plants, but are accumulated by soil
macroinvertebrates. Thus, in areas without significant soil deposition on the surfaces of plants,
mammalian herbivores would not be exposed to PCBs in most of their diet. In contrast,
mammalian insectivores, such as shrews, could be exposed to PCBs in most of their diet. For
PCBs, the ingestion route for mammals would be essentially incomplete for herbivores but
complete for insectivores. For the PCB site, therefore, the ingestion exposure route for a
mammalian insectivore (e.g., shrew) would be a complete exposure pathway that should be
evaluated.
Incidental soil/sediment ingestion. Mammals can ingest some quantity of soils or
sediments incidentally, as they groom their fur or consume plants or animals from the soil.
Burrowing mammals are likely to ingest greater quantities of soils during grooming than non-
burrowing mammals, and mammals that consume plant roots or soil-dwelling macroinvertebrates
are likely to ingest greater quantities of soils attached to the surface of their foods than
mammals that consume other foods. The intake of PCBs from incidental ingestion of PCB-
contaminated soils is difficult to estimate, but for insectivores that forage at ground level, it is
likely to be far less than the intake of PCBs in the diet. For herbivores, the incidental intake of
PCBs in soils might be higher than the intake of PCBs in their diet, but still less than the intake
of PCBs by mammals feeding on soil macroinvertebrates. Thus, the exposure pathway for
ground-dwelling mammalian insectivores remains the exposure pathway that should be
evaluated.
1.3 SCREENING-LEVEL ECOLOGICAL EFFECTS EVALUATION
The next step in the screening-level risk assessment is the preliminary ecological
effects evaluation and the establishment of contaminant exposure levels that represent
conservative thresholds for adverse ecological effects. In this guidance, those conservative
thresholds are called screening ecotoxicity values. Physical stresses unrelated to contaminants
at the site are not the focus of the risk assessment (see Highlight 1-4), although they can be
considered later when evaluating effects of remedial alternatives.
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A literature search for studies that
quantify toxicity (i.e., exposure-response) is
necessary to evaluate the likelihood of toxic
effects in different groups of organisms.
Appendix C provides a basic introduction to
conducting a literature search, but an expert
should be consulted to minimize time and
costs. The toxicity profile should describe
the toxic mechanisms of action for the
exposure routes being evaluated and the
dose or environmental concentration that
causes a specified adverse effect.
For each complete exposure pathway,
route, and contaminant, a screening
ecotoxicity value should be developed.1
The U.S. EPA Office of Emergency and
Remedial Response has developed screening
ecotoxicity values [called ecotox threshold
values (U.S. EPA, 1996c)]. The values are
for surface waters and sediments, and are
based on direct exposures routes only;
bioaccumulation and biomagnification in
food chains have not been accounted for.
The following subsections describe preferred
data (Section 1.3.1), dose conversions
(Section 1.3.2), and analyzing uncertainty in
the values (Section 1.3.3).
1.3.1 Preferred Toxicity Data
Screening ecotoxicity values should represent a no-observed-adverse-effect-level
(NOAEL) for long-term (chronic) exposures to a contaminant. Ecological effects of most
concern are those that can impact populations (or higher levels of biological organization).
Those include adverse effects on development, reproduction, and survivorship. Community-
level effects also can be of concern, but toxicity data on community-level endpoints are
limited and might be difficult to extrapolate from one community to another.
HIGHLIGHT 1-4
Non-Chemical Stressors
Ecosystems can be stressed by
physical, as well as by chemical, alterations
of their environment. For this reason,
EPA's (1992a) Framework for Ecological
Risk Assessment addresses "stressor-
response" evaluation to include all types of
stress instead of "dose-response" or
"exposure-response" evaluation, which
implies that the stressor must be a toxic
substance.
For Superfund sites, however, the
baseline risk assessment addresses risks from
hazardous substances released to the
environment, not risks from physical
alterations of the environment, unless caused
indirectly by a hazardous substances (e.g.,
loss of vegetation from a chemical release
leading to serious erosion). This guidance
document, therefore, focuses on exposure-
response evaluations for toxic substances.
Physical destruction of habitat that might be
associated with a particular remedy is
considered in the Feasibility Study.
' It is possible to conduct a screening risk assessment with limited information and conservative
assumptions. If site-specific information is too limited, however, the risk assessment is almost certain to move
into Steps 3 through 7. which require field-collected data. The more complete the initial information, the better
the decision that can be made at this preliminary stage.
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When reviewing the literature, one
should be aware of the limitations of
published information in characterizing
actual or probable hazards at a specific site.
U.S. EPA discourages reliance on secondary
references because study details relevant for
determining the applicability of findings to a
given site usually are not reported in
secondary sources. Only primary literature
that has been carefully reviewed by an
ecotoxicologist should be used to support a
decision. Several considerations and data
preferences are summarized ,in Highlight 1-5
and described more fully below.
NOAELS and LOAELS. For each
contaminant for which a complete exposure
pathway/route exists, the literature should be
reviewed for the lowest exposure level (e.g.,
concentration in water or in the diet, ingested
dose) shown to produce adverse effects (e.g.,
reduced growth, impaired reproduction,
increased mortality) in a potential receptor
species. This value is called a lowest-
observed-adverse-effect-level or LOAEL.
For those contaminants with documented
adverse effects, one also should identify the
highest exposure level that is a NOAEL. A
NOAEL is more appropriate than a LOAEL
to use as an screening ecotoxicity value to
ensure that risk is not underestimated (see Highlight 1-6). However, NOAELs currently are
not available for many groups of organisms and many chemicals. When a LOAEL value, but
not a NOAEL value, is available from the literature, a standard practice is to multiply the
LOAEL by 0.1 and to use the product as the screening ecotoxicity value. Support for this
practice comes from a data review indicating that 96 percent of chemicals included in the
review had LOAEL/NOAEL ratios of five or less, and that all were ten or less (Dourson and
Stara, 1983).
Exposure duration. Data from studies of chronic exposure are preferable to data
from medium-term (subchronic), short-term (acute), or single-exposure studies because
exposures at Superfund remedial sites usually are long-term. Literature reviews by
McNamara (1976) and Weil and McCollister (1963) indicate that chronic NOAELs can be
HIGHLIGHT 1-5
Data Hierarchy for Deriving
Screening Ecotoxicity Values
To develop a chronic NOAEL for a
screening ecotoxicity value from existing
literature, the following data hierarchy
minimizes extrapolations and uncertainties
in the value:
• A NOAEL is preferred to a
LOAEL, which is preferred to an
LC50 or an EC50.
• Long-term (chronic) studies are
preferred to medium-term
(subchronic) studies, which are
preferred to short-term (acute)
studies.
• If exposure at the site is by
ingestion, dietary studies are
preferred to gavage studies, which
are preferred to non-ingestion routes
of exposure. Similarly, if exposure
at the site is dermal, dermal studies
are preferred to studies using other
exposure routes.
MO
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lower than subchronic (90-day duration for
rats) NOAELs by up to a factor of ten.2
N . Exposure route. The exposure
route and medium used in the toxicity study
should be comparable to the exposure route
in the risk assessment. For example, data
from studies where exposure is by gavage
generally are not preferred for estimating
dietary concentrations that could produce
adverse effects, because the rate at which
the substance is absorbed from the
gastrointestinal tract usually is greater
following gavage than following dietary
administration. Similarly, intravenous
injection of a substance results in
"instantaneous absorption" and does not
allow the substance to first pass through the
liver, as it would following dietary
exposure. If it is necessary to attempt to
extrapolate toxicity test results from one
route of exposure to another, the
extrapolation should be performed or
reviewed by a toxicologist experienced in
route-to-route extrapolations for the class of
HIGHLIGHT 1-6
NOAEL Preferred to LOAEL
Because the NOAEL and LOAEL
are estimated by hypothesis testing (i.e., by
comparing the response level of a test group
to the response level of a control group for a
statistically significant difference), the actual
proportion of the test animals showing the
adverse response at an identified LOAEL
depends on sample size, variability of the
response, and the dose interval. LOAELs,
and even NOAELs, can represent a
30 percent or higher effect level for the
minimum sample sizes recommended for
standard test protocols. For this reason, U.S.
EPA recommends that the more conservative
NOAELs, instead of LOAELs, are used to
determine a .screening exposure level that is.
unlikely to adversely impact populations. If
dose-response data are available, a site-
specific low-effect level may be determined.
animals at issue.
Field versus laboratory. Most toxicity studies evaluate effects of a single
contaminant on a single species under controlled laboratory conditions. Results from these
studies might not be directly applicable to the field, where organisms typically are exposed to
more than one contaminant in environmental situations that are not comparable to a laboratory
setting and where genetic composition of the population can be more heterogeneous than that
of organisms bred for laboratory use. In addition, the bioavailability of a contaminant might
be different at a site than in a laboratory toxicity test. In a field situation, organisms also will
be subject to other environmental variables, such as unusual weather conditions, infectious
diseases, and food shortages. These variables can have either positive or negative effects on
2 The literature reviews of McNamara (1976) and Weil and McCollister (1963) included both rodent and
non-rodent species. The duration of the subchronic exposure usually was 90 days, but ranged from 30 to 210
days. A wide variety of endpoints and criteria for adverse effects were included in these reviews. Despite this
variation in the original studies, their findings provide a general indication of the ratio between subchronic to
chronic NOAELs for effects other than cancer and reproductive effects. For some chemicals, chronic dosing
resulted in increased chemical tolerance. For over 50 percent of the compounds tested, the chronic NOAEL was
less than the 90-day NOAEL by a factor of 2 or less. However, in a few cases, the chronic NOAEL was up to a
factor of 10 less than the subchronic NOAEL (U.S. EPA, 1993e).
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the organism's response to a toxic contaminant that only a site-specific field study would be
able to evaluate. Moreover, single-species toxicity tests seldom provide information regarding
toxicant-related changes in community interactions (e.g., behavioral changes in prey species
that make them more susceptible to predation).
1.3.2 Dose Conversions
For some data reported in the literature, conversions are necessary to allow the data to
be used for species other than those tested or for measures of exposure other than those
reported. Many doses in laboratory studies are reported in terms of concentration in the diet
(e.g., mg contaminant/kg diet or ppm in the diet). Dietary concentrations can be converted to
dose (e.g., mg contaminant/kg body weight/day) for comparison with estimated contaminant
intake levels in the receptor species.
When converting doses, it is important to identify whether weights are measured as
wet or dry weights. Usually, body weights are reported on a wet-weight, not dry-weight
basis. Concentration of the contaminant in the diet might be reported on a wet- or dry-weight
basis.
Ingestion rates and body weights for a test species often are reported in a toxicity
study or can be obtained from other literature sources (e.g., U.S. EPA, 1993a,b). For
extrapolations between animal species with different metabolic rates as well as dietary
composition, consult U.S. EPA 1992e and 1996b.
1.3.3 Uncertainty Assessment
Professional judgment is needed to determine the uncertainty associated with
information taken from the literature and any extrapolations used in developing a screening
ecotoxicity value. The risk assessor should be consistently conservative in selecting literature
values and describe the limitations of using those values in the context of a particular site.
Consideration of the study design, endpoints, and other factors are important in determining
the utility of toxicity data in the screening-level risk assessment. All of those factors should
be addressed in a brief evaluation of uncertainties prior to the screening-level risk calculation.
1.4 SUMMARY
At the conclusion of the screening-level problem formulation and ecological effects
evaluation, the following information should have been compiled:
• Environmental setting and contaminants known or suspected to exist at the site
and the maximum concentrations present (for each medium);
• Contaminant fate and transport mechanisms that might exist at the site;
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• The mechanisms of ecotoxicity associated with contaminants and likely
categories of receptors that could be affected;
• The complete exposure pathways that might exist at the site from contaminant
sources to receptors that could be affected; and
• Screening ecotoxicity values equivalent to chronic NOAELs based on
conservative assumptions.
For the screening-level ecological risk assessment, assessment endpoints will include
any likely adverse ecological effects on receptors for which exposure pathways are complete,
as determined from the information listed above. Measurement endpoints will be based on
the available literature regarding mechanisms of toxicity and will be used to establish the
screening ecotoxicity values. Those values will be used with estimated exposure levels to
screen for ecological risks, as described in Step 2.
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STEP 2: SCREENING-LEVEL EXPOSURE ESTIMATE
AND RISK CALCULATION
OVERVIEW
The screening-level exposure estimate and risk calculation comprise the second
step in the ecological risk screening for a site. Risk is estimated by comparing
maximum documented exposure concentrations with the ecotoxicity screening values
from Step 1. At the conclusion of Step 2, the risk manager and risk assessment team
will decide that either the screening-level ecological risk assessment is adequate to
determine that ecological threats are negligible, or the process should continue to a
more detailed ecological risk assessment (Steps 3 through 7). If the process continues,
the screening-level assessment serves to identify exposure pathways and preliminary
contaminants of concern for the baseline risk assessment by eliminating those
contaminants and exposure pathways that pose negligible risks.
2.1 INTRODUCTION
This step includes estimating exposure levels and screening for ecological risks as the
last two phases of the screening-level ecological risk assessment. The process concludes with
a SMDP at which it is determined that: (1) ecological threats are negligible; (2) the
ecological risk assessment should continue to.determine whether a risk exists; or (3) there is a
potential for adverse ecological effects, and a more detailed ecological risk assessment,
incorporating more site-specific information, is needed.
Section 2.2 describes the screening-level exposure assessment, focusing on the
complete exposure pathways identified in Step 1. Section 2.3 describes the risk calculation
process, including estimating a hazard quotient, documenting the uncertainties in the quotient,
and summarizing the overall confidence in the screening-level ecological risk assessment.
Section 2.4 describes the SMDP that concludes Step 2.
2.2 SCREENING-LEVEL EXPOSURE ESTIMATES
To estimate exposures for the screening-level ecological risk calculation, on-site
contaminant levels and general information on the types of biological receptors that might be
exposed should be known from Step 1. Only complete exposure pathways should be
evaluated. For these, the highest measured or estimated on-site contaminant concentration for
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each environmental medium should be used to estimate exposures. This should ensure that
potential ecological threats are not missed.
2.2.1 Exposure Parameters
For parameters needed to estimate exposures for which sound site-specific information
is lacking or difficult to develop, conservative assumptions should be used at this screening
level. Examples of conservative assumptions are listed below and described in the following
paragraphs:
• Area-use factor - 100 percent (factor
related to home range and population
density; see Highlight 2-1);
• Unavailability - 100 percent;
• Life stage - most sensitive life stage;
• Body weight and food ingestion rate
- minimum body weight to
maximum ingestion rate; and
• Dietary composition - 100 percent of
diet consists of the most
contaminated dietary component.
Area-use factor. For the
screening-level exposure estimate for
terrestrial animals, assume that the home
range of one or more animals is entirely within the contaminated area, and thus the animals
are exposed 100 percent of the time. This is a conservative assumption and, as an
assumption, is only applicable to the screening-level phase of the risk assessment. Species-
and site-specific home range information would be needed later, in Step 6, to estimate more
accurately the percentage of time an animal would use a contaminated area. Also evaluate
the possibility that some species might actually focus their activities in contaminated areas of
the site. For example, if contamination has reduced emergent vegetation in a pond, the pond
might be more heavily used for feeding by waterfowl than uncontaminated ponds with little
open water.
Bioavailability. For the screening-level exposure estimate, in the absence of site-
specific information, assume that the bioavailability of contaminants at the site is 100 percent.
For example, at the screening-level, lead would be assumed to be 100 percent bioavailable to
mammals. While some literature indicates that mammals absorb approximately 10 percent of
ingested lead, absorption efficiency can be higher, up to about 60 percent, because dietary
HIGHLIGHT 2-1
Area-use Factor
An animal's area-use factor can be
defined as the ratio of the area of
contamination (or the site area under
investigation) to the area' used by the animal,
e.g., its home range, breeding range, or
feeding/foraging range. To ensure that
ecological risks are not underestimated, the
highest density and smallest area used by
each animal should be assumed. This allows
the maximum number of animals to be
exposed to site contaminants and makes it
more likely that "hot spots" (i.e., areas of
unusually high contamination levels) will be
significant proportions of an individual
animal's home range.
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factors such as fasting, and calcium and phosphate content of the diet, can affect the
absorption rate (Kenzaburo, 1986). Because few species have been tested for bioavailability,
and because Steps 3 through 6 provide an opportunity for this issue to be addressed
specifically, the most conservative assumption is appropriate for this step.
Life stage. For the screening-level assessment, assume that the most sensitive life
stages are present. If an early life stage is the most sensitive, the population should be
assumed to include or to be in that life stage. For vertebrate populations, it is likely that most
of the population is not in the most sensitive life stage most of the time. However, for many
invertebrate species, the entire population can be at an early stage of development during
certain seasons.
Body weight and food ingestion rates. Estimates of body weight and food
ingestion rates of the receptor animals also should be made conservatively to maximize the
dose (intake of contaminants) on a body-weight basis and to avoid understating risk, although
uncertainties in these factors are far less than the uncertainties associated with the
environmental contaminant concentrations. U.S. EPA's Wildlife Exposure Factors Handbook
(U.S. EPA, 1993a,b) is a good source or reference to sources of this information.
Bioaccumulation. Bioaccumulation values obtained from a literature search can be
used to estimate contaminant accumulation and food-chain transfer at a Superfund site at the
screening stage. Because many environmental factors influence the degree of
bioaccumulation, sometimes by several orders of magnitude, the most conservative (i.e.,
highest) bioaccumulation factor (BAF) reported in the literature should be used in the absence
of site-specific information.
Dietary composition. For species that feed on more than one type of food, the
screening-level assumption should be that the diet is composed entirely of whichever type of
food is most contaminated. For example, if some foods (e.g., insects) are likely to be more
contaminated than other foods (e.g., seeds and fruits) typical in the diet of a receptor species,
assume that the receptor species feeds exclusively on the more contaminated type of food.
Again, EPA's Wildlife Exposure Factors Handbook (U.S. EPA, 1993a,b) is a good source or
reference to sources of this information.
2.2.2 Uncertainty Assessment
Professional judgment is needed to determine the uncertainty associated with
information taken from the literature and any extrapolations used in developing a parameter to
estimate exposures. All assumptions used to estimate exposures should be stated, including
some description of the degree of bias possible in each. Where literature values are used, an
indication of the range of values that could be considered appropriate also should be
indicated.
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2.3 SCREENING-LEVEL RISK CALCULATION
A quantitative screening-level risk can be estimated using the exposure estimates
developed according to Section 2.2 and the screening ecotoxicity values developed according
to Section 1.3. For the screening-level risk calculation, the hazard quotient approach, which
compares point estimates of screening ecotoxicity values and exposure values, is adequate to
estimate risk. As described in Section 1.3, a screening ecotoxicity value should be equivalent
to a documented and/or best conservatively estimated chronic NOAEL. Thus, for each
contaminant and environmental medium, the hazard quotient can be expressed as the ratio of
a potential exposure level to the NOAEL:
BQ . or HQ
NOAEL NOAEL
where: .
HQ = hazard quotient;
Dose = estimated contaminant intake at the site (e.g., mg contaminant/kg body
weight per day);
EEC = estimated environmental concentration at the site (e.g., mg
contarninant/L water, mg contaminant/kg soil, mg contaminant/kg food);
and
NOAEL = no-observed-adverse-effects-level (in units that match the dose or EEC).
An HQ less than one (unity) indicates that the contaminant alone is unlikely to cause adverse
ecological effects. If multiple contaminants of potential ecological concern exist at the site, it
might be appropriate to sum the HQs for receptors that could be simultaneously exposed to
the contaminants that produce effects by the same toxic mechanism (U.S. EPA, 1986a). The
sum of the HQs is called a hazard index (HI); (see Highlight 2-2). An HI less than one
indicates that the group of contaminants is unlikely to cause adverse ecological effects. An
HQ or HI less than one does not indicate the absence of ecological risk; rather, it should be
interpreted based on the severity of the effect reported and the magnitude of the calculated
quotient. As certainty in the exposure concentrations and the NOAEL increase, there is
greater confidence in the predictive value of the hazard quotient model, and unity (HQ = 1)
becomes a more certain pass/fail decision point.
The screening-level risk calculation is a conservative estimate to ensure that potential
ecological threats are not overlooked. The calculation is used to document a decision about
whether or not there is a negligible potential for ecological impacts, based on the information
available at this stage. If the potential for ecological impacts exists, this calculation can be
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used to eliminate the negligible-risk
combinations of contaminants and exposure
pathways from further consideration.
If the screening-level risk assessment
indicates that adverse ecological effects are
possible at environmental concentrations
below standard quantitation limits, a "non
detect" based on those limits cannot be used
to support a "no risk" decision. Instead, the
risk assessment team and risk manager
should request appropriate detection limits
or agree to continue to Steps 3 through 7,
where exposure concentrations will be
estimated from other information (e.g., fate-
and-transport modeling, assumed or
estimated values for non-detects).
2.4 SCIENTIFIC/MANAGEMENT
DECISION POINT (SMDP)
HIGHLIGHT 2-2
Hazard Index (HI) Calculation
For contaminants that produce adverse
effects by the same toxic mechanism:
Hazard Index =
EEC,/NOAEL,
EEC2/NOAEL2
EECj/NOAELj
where:
estimated environmental
concentration for the 1th
contaminant; and
NOAELi =
NOAEL for the i* contaminant
(expressed either as a dose or
environmental concentration).
The EEC and the NOAEL are expressed in
the same units and represent the same
exposure period (e.g., chronic). Dose could
be substimted for EEC throughout provided
the NOAEL is expressed as a dose.
At the end of Step 2, the lead risk
assessor communicates the results of the
preliminary ecological risk assessment to the
risk manager. The risk manager needs to
decide whether the information available is
adequate to make a risk management
decision and might require technical advice from the ecological risk assessment team to reach
a decision. There are only three possible decisions at this point:
(1) There is adequate information to conclude that ecological risks are negligible
and therefore no need for remediation on the basis of ecological risk;
(2) The information is not adequate to make a decision at this point, and the
ecological risk assessment process will continue to Step 3; or
(3) The information indicates a potential for adverse ecological effects, and a more
thorough assessment is warranted.
Note that the SMDP made at the end of the screening-level risk calculation will not
set a preliminary cleanup goal. Screening ecotoxicity values are derived to avoid
underestimating risk. Requiring a cleanup based solely on those values would not be
technically defensible.
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The risk manager should document both the decision and the basis for it. If the risk
characterization supports the first decision (i.e., negligible risk), the ecological risk assessment
process ends here with appropriate documentation to support the decision. The documentation
should include all analyses and references used in the assessment, including a discussion of
the uncertainties associated with the HQ and HI estimates.
For assessments that proceed to Step 3, the screening-level analysis in Step 2 can
indicate and justify which contaminants and exposure pathways can be eliminated from
further assessment because they are unlikely to pose a substantive risk, (If new contaminants
are discovered or contaminants are found at higher concentrations later in the site
investigation, those contaminants might need to be added to the ecological risk assessment at
that time.)
U.S. EPA must be confident that the SMDP made after completion of this calculation
will protect the ecological components of the environment. The decision to continue beyond
the screening-level risk calculation does not indicate whether remediation is necessary at the
site. That decision will be made in Step 8 of the process.
2.5 SUMMARY
At the conclusion of the exposure estimate and screening-level risk calculation step,
the following information should have been compiled:
(1) Exposure estimates based on conservative assumptions and maximum
concentrations present; and
(2) Hazard quotients (or hazard indices) indicating which, if any, contaminants and
exposure pathways might pose ecological threats.
Based on the results of the screening-level ecological risk calculation, the risk manager
and lead risk assessor will determine whether or not contaminants from the site pose an
ecological threat. If there are sufficient data to determine that ecological threats are
negligible, the ecological risk assessment will be complete at this step with a finding of
negligible ecological risk. If the data indicate that there is (or might be) a risk of adverse
ecological effects, the ecological risk assessment process will continue.
(
Conservative assumptions have been used for each step of the screening-level
ecological risk assessment. Therefore, requiring a cleanup based solely on this information
would not be technically defensible. To end the assessment at this stage, the conclusion of
negligible ecological risk must be adequately documented and technically defensible. A lack
of information on the toxicity of a contaminant or on complete exposure pathways will result
in a decision to continue with the ecological risk assessment process (Steps 3 through 7)—not
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a decision to delay the ecological risk assessment until a later date, when more information
might be available.
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STEP 3: BASELINE RISK ASSESSMENT PROBLEM FORMULATION
OVERVIEW
Step 3 of the eight-step process initiates the problem-formulation phase of the
baseline ecological risk assessment. Step 3 refines the screening-level problem
formulation and, with input from stakeholders and other involved parties, expands on
the ecological issues that are of concern at the particular site. In the screening-level
assessment, conservative assumptions were used where site-specific information was
lacking. In Step 3, the results of the screening assessment and additional site-specific
information are used to determine the scope and goals of the baseline ecological risk
assessment. Steps 3 through 7 are required only for sites for which the screening-level
assessment indicated a need for further ecological risk evaluation.
Problem formulation at Step 3 includes several activities:
• Refining preliminary contaminants of ecological concern;
• Further characterizing ecological effects of contaminants;
• Reviewing and refining information on contaminant fate and transport, complete
exposure pathways, and ecosystems potentially at risk;
• Selecting assessment endpoints; and
• Developing a conceptual model with working hypotheses or questions that the
site investigation will address.
At the conclusion of Step 3, there is a SMDP, which consists of agreement on four
items: the assessment endpoints, the exposure pathways, the risk questions, and
conceptual model integrating these components. The products of Step 3 are used to
select measurement endpoints and to develop the ecological risk assessment work plan
(WP) and sampling and analysis plan (SAP) for the site in Step 4. Steps 3 and 4 are,
effectively, the data quality objective (DQO) process for the baseline ecological risk
assessment.
3.1 THE PROBLEM-FORMULATION PROCESS
In Step 3, problem formulation establishes the goals, breadth, and focus of the baseline
ecological risk assessment. It also establishes the assessment endpoints, or specific ecological
values to be protected (U.S. EPA, 1992a). Through Step 3, the questions and issues that need
to be addressed in the baseline ecological risk assessment are defined based on potentially
complete exposure pathways and ecological effects. A conceptual model of the site is
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developed that includes questions about the assessment endpoints and the relationship between
exposure and effects. Step 3 culminates in an SMDP, which is agreement between the risk
manager and risk assessor on the assessment endpoints, exposure pathways, and questions as
portrayed in the conceptual model of the site.
The conceptual model, which is completed in Step 4, also will describe the approach,
types of data, and analytical tools to be used for the analysis phase of the ecological risk
assessment (Step 6). Those components of the conceptual model are formally described in
the ecological risk WP and SAP in Step 4 of this eight-step process. If there is not
agreement among the risk manager, lead risk assessor, and the other professionals involved
with the ecological risk assessment on the initial conceptual model developed in Step 3, the
final conceptual model and field study design developed in Step 4 might not resolve the
issues that must be considered to manage risks effectively.
The complexity of questions developed during problem formulation does not depend
on the size of a site or the magnitude of its contamination. Large areas of contamination can
provoke simple questions and, conversely, small sites with numerous contaminants can require
a complex series of questions and assessment endpoints. There is no rule that can be applied
to gauge the effort needed for an ecological risk assessment based on site size or number of
contaminants; each site should be evaluated individually.
At the beginning of Step 3, some basic information should exist for the site. At a
minimum, information should be available from the site history, PA, SI, and Steps 1 and 2 of
this eight-step process. For large or complex sites, information might be available from
earlier site investigations.
It is important to be as complete as possible early in the process so that Steps 3
through 8 need not be repeated. Repeating the selection of assessment endpoints and/or the
questions and hypotheses concerning those endpoints is appropriate only if new information
indicating new threats becomes available. The SMDP process should prevent having to return
to the problem formulation step because of changing opinions on the questions being asked.
Repetition of Step 3 should not be confused with the intentional tiering (or phasing) of
ecological site investigations at large or complex sites (see Highlight 3-1). The process of
problem formulation at complex sites is the same as at more simple sites, but the number,
complexity, and/or level of resolution of the questions and hypotheses can be greater at
complex sites.
While problem formulation is conceptually simple, in practice it can be a complex and
interactive process. Defining the ecological problems to be addressed during the baseline risk
assessment involves identifying toxic mechanisms of the contaminants, characterizing
potential receptors, and estimating exposure and potential ecological effects. Problem
formulation also constitutes the DQO process for the baseline ecological risk assessment (U.S.
EPA, 1993c,d).
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The remainder of this section
describes six activities to be conducted
prior to the SMDP for this step:
refining preliminary contaminants of
ecological concern (Section 3.2); a
literature search on the potential
ecological effects of the contaminants
(Section 3.3); qualitative evaluation of
complete exposure pathways and
ecosystems potentially at risk (Section
3.4); selecting assessment endpoints
(Section 3.5); and developing the
conceptual model and establishing risk
questions (Section 3.6).
HIGHLIGHT 3-1
Tiering an Ecological Risk
Assessment
Most ecological risk assessments at
Superfund sites are at least a two-tier process.
Steps 1 and 2 of this guidance serve as a first,
or screening, tier prior to expending a larger
effort for a detailed, site-specific ecological risk
assessment. The baseline risk assessment may
serve as the second tier. Additional tiers could
be needed in the baseline risk assessment for
large or complex sites where there is a need to
sequentially test interdependent hypotheses
developed during problem formulation (i.e.,
evaluating the results of one field assessment
before designing a subsequent field study).
While tiering can be an effective way to
manage site investigations, multiple sampling
phases typically require some resampling of
matrices sampled during earlier tiers and
increased field-mobilization costs. Thus, in
some cases, a multi-tiered ecological risk
assessment might cost more than a two-tiered
assessment. The benefits of tiering should be
weighed against the costs.
3.2 REFINEMENT OF
PRELIMINARY
CONTAMINANTS OF
CONCERN
The results of the screening-level
risk assessment (Steps 1 and 2) should
have indicated which contaminants
found at the site can be eliminated from
further consideration and which should
be evaluated further. It is important to
realize that contaminants that might pose
an ecological risk can be different from
those that might pose a human health risk because of differing exposure pathways,
sensitivities, and responses to contaminants.
The initial list of contaminants investigated in Steps 1 and 2 included all contaminants
identified or suspected to be at the site. During Steps 1 and 2, it is likely that several of the
contaminants found at the site were eliminated from further assessment because the risk
screen indicated that they posed a negligible ecological risk. Because of the conservative
assumptions used during the risk screen, some of the contaminants retained for Step 3 might
also pose negligible risk. At this stage, the risk assessor should review the assumptions used
(e.g., 100 percent bioavailability) against values reported in the literature (e.g., only up to 60
percent for a particular contaminant), and consider how the HQs would change if more
realistic conservative assumptions were used instead (see Section 3.4.1). For those
contaminants for which the HQs drop to near or below unity, the lead risk assessor and risk
manager should discuss and agree on which can be eliminated from further consideration at
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this time. The reasons for dropping any contaminants from consideration at this step must be
documented in the baseline risk assessment.
Sometimes, new information becomes available that indicates the initial assumptions
that screened some contaminants out in Step 2 are no longer valid (e.g., site contaminant
levels are higher than originally reported). In this case, contaminants can be placed back on
the list of contaminants to be investigated with that justification.
Note that a contaminant should not be eliminated from the list of contaminants to be
investigated only because toxicity information is lacking; instead, limited or missing toxicity
information must be addressed using best professional judgment and discussed as an
uncertainty.
3.3 LITERATURE SEARCH ON KNOWN ECOLOGICAL EFFECTS
The literature search conducted in Step 1 for the screening-level risk assessment might
need to be expanded to obtain the information needed for the more detailed problem
formulation phase of the baseline ecological risk assessment. The literature search should
identify NOAELs, LOAELs, exposure-response functions, and the mechanisms of toxic
responses for contaminants for which those data were not collected in Step 1. Appendix C
presents a discussion of some of the factors important in conducting a literature search.
Several U.S. EPA publications (e.g., U.S. EPA, 1995a,e,g,h) provide a window to original
toxicity literature for contaminants often found at Superfund sites. For all retained
contaminants, it is important to obtain and review the primary literature.
3.4 CONTAMINANT FATE AND TRANSPORT, ECOSYSTEMS POTENTIALLY AT
RISK, AND COMPLETE EXPOSURE PATHWAYS
A preliminary identification of contaminant fate and transport, ecosystems potentially
at risk, and complete exposure pathways was conducted in the screening ecological risk
assessment. In Step 3, the exposure pathways and the ecosystems associated with the
assessment endpoints that were retained by the screening risk assessment are evaluated in
more detail. This effort typically involves compiling additional information on:
(1) The environmental fate and transport of the contaminants;
(2) The ecological setting and general flora and fauna of the site (including habitat,
potential receptors, etc.); and
(3) The magnitude and extent of contamination, including its spatial and temporal
variability relative to the assessment endpoints.
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For individual contaminants, it is frequently possible to reduce the number of exposure
pathways that need to be evaluated to one or a few "critical exposure pathways" which (1)
reflect maximum exposures of receptors within the ecosystem, or (2) constitute exposure
pathways to ecological receptors sensitive to the contaminant. The critical exposure pathways
influence the selection of assessment endpoints for a particular site. If multiple critical
exposure pathways exist, they each should be evaluated, because it is often difficult to predict
which pathways could be responsible for the greatest ecological risk.
3.4.1 Contaminant Fate and Transport
Information on how the contaminants will or could be transported or transformed in
the environment physically, chemically, and biologically is used to identify the exposure
pathways that might lead to significant ecological effects (see Highlight 3-2). Chemically,
contaminants can undergo several processes in the environment:
• Degradation,3
• Complexation,
• lonization,
• Precipitation, and/or
• Adsorption.
Physically, contaminants might move
through the environment by one or more
means:
• Volatilization,
• Erosion,
• Deposition (contaminant
sinks),
• Weathering of parent material
with subsequent transport,
and/or
• Water transport:
in solution,
as suspended material in the water, and
bulk transport of solid material.
Several biological processes also affect contaminant fate and transport in the environment:
• Bioaccumulation,
• Biodegradation,
HIGHLIGHT 3-2
Environmental Fate and Exposure
If a contaminant in an aquatic
ecosystem is highly lipophilic (i.e.,
essentially insoluble in water), it is likely to
partition primarily into sediments and not
into the water column. Factors such as
sediment particle size and organic carbon
influence contaminant partitioning; therefore,
these attributes should be characterized when
sampling sediments. Similar considerations
regarding partitioning should be applied to
contaminants in soils.
The product might be more or less toxic than the parent compound.
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• Biological transformation,4
• Food chain transfers, and/or
• Excretion.
Additional information should be gathered on past as well as current mechanisms of
contaminant release from source areas at the site. The mechanisms of release along with the
chemical and physical form of a contaminant can affect its fate, transport, and potential for
reaching ecological receptors.
A contaminant flow diagram (or exposure pathway diagram) comprises a large part of
the conceptual model, as illustrated in Section 3.6. A contaminant flow diagram originates at
the primary contaminant source(s) and identifies primary release mechanisms and contaminant
transport pathways. The release and movement of the contaminants can create secondary
sources (e.g., contaminated sediments in a river; see Example 3-1), and even tertiary sources.
The above information is used to evaluate where the contaminants are likely to
partition in the environment, and the bioavailability of the contaminant (historically, currently,
or in the future). As indicated in Section 3.2, it might be possible for the risk assessment
team and the risk manager to use this information to replace some of the conservative
assumptions used in the screening-level risk assessment and to eliminate additional chemicals
from further evaluation at this point. Any such negotiations must be documented in the
baseline risk assessment.,
3.4.2 Ecosystems Potentially at Risk
The ecosystems or habitats potentially at risk depend oh the ecological setting of a
site. An initial source of information on the ecological setting of a site is the data collected
during the preliminary site visit and characterization (Step 1), including the site ecological
checklist (Appendix B). The site description should provide answers to several questions
including:
• What habitats (e.g., maple-beech hardwood forest, early-successional fields) are
present?
• What types of water bodies are present, if any?
• Do any other habitats listed in Exhibit 1-1 exist on or adjacent to the site?
While adequately documented information should be used, it is not critical that
complete site setting information be collected during this phase of the risk assessment.
However, it is important that habitats at the site are not overlooked; hence, a site visit might
be needed to supplement the one conducted during the screening risk assessment. If a habitat
The product might be more or less toxic than the parent compound.
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EXAMPLE 3-1
Exposure Pathway Model-DDT Site
An abandoned pesticide production facility had released DDT to soils through poor
handling practices during its operation. Due to erosion of contaminated soils, DDT migrated to
stream sediments. The contaminated sediments represent a secondary source that might affect
benthic organisms through direct contact or ingestion. Benthic organisms that have accumulated
DDT can be consumed by fish, and fish that have accumulated DDT can be consumed by
piscivorous birds, which are considered a valuable component of the local ecosystem. This
example illustrates how contaminant transport is traced from a primary source to a secondary
source and from there through a food chain to an exposure point that can affect an assessment
endpoint.
actually present on the site is omitted during the problem formulation phase, this step might
need to be repeated later when the habitat is found, resulting in delays and additional costs
for the risk assessment.
Available information on ecological effects of contaminants (see Section 3.3) can help
focus the assessment on specific ecological resources that should be evaluated more
thoroughly, because some groups of organisms can be more sensitive than others to a
particular contaminant. For example, a species or group of species could be physiologically
sensitive to a particular contaminant (e.g., the contaminant might interfere with its vascular
system); or, the species might not be able to metabolize and detoxify the particular
contaminant(s) (e.g., honey bees and grass shrimp cannot effectively biodegrade PAHs,
whereas fish generally can). Alternatively, an already-stressed population (e.g., due to habitat
degradation) could be particularly sensitive to any added stresses.
Variation in sensitivity should not be confused with variation in exposure, which can
result from behavioral and dietary differences among species. For example, predators can be
exposed to higher levels of contaminants that biomagnify in food chains than herbivores. A
specialist predator could feed primarily on one prey type that is a primary receptor of the
contaminant. Some species might preferentially feed in a habitat where the contaminant tends
to accumulate. On the other hand, a species might change its behavior to avoid contaminated
areas. Both sensitivity to toxic effects of a contaminant and behaviors that affect exposure
levels can influence risks for particular groups of organisms.
3.4.3 Complete Exposure Pathways
The potentially complete exposure pathways identified in Steps 1 and 2 are described
in more detail in Step 3 on the basis of the refined contaminant fate and transport evaluations
(Section 3.4.1) and evaluation of potential ecological receptors (Section 3.4.2).'
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Some of the potentially complete exposure pathways identified in Steps 1 and 2 might
be ruled out from further consideration at this time. Sometimes, additional exposure
pathways might be identified, particularly those originating from secondary sources. Any data
gaps that result in questions about whether an exposure pathway is complete should be
identified, and the type of data needed to answer those questions should be described to assist
in developing the WP and SAP in Step 4. .
During Step 3, the potential for food-chain exposures deserves particular attention.
Some contaminants are effectively transferred through food chains, while others are not. To
illustrate this point, copper and DDT are compared in Example 3-2.
EXAMPLE 3-2
Potential for Food Chain Transfer-Copper and DDT Sites
Copper can be toxic in aquatic ecosystems and to terrestrial plants. However, it is an
essential nutrient for both plants and animals, and organisms can regulate internal copper
concentrations within limits. For this reason, copper tends not to accumulate in most organisms
or to biomagnify in food chains, and thus tends not to reach levels high enough to cause
adverse responses through food chain transfer to upper-trophic-level organisms. (Copper is
known to accumulate by several orders of magnitude in phytoplankton and in filter-feeding
mollusks, however, and thus can pose a threat to organisms that feed on those components of
aquatic ecosystems; U.S. EPA, 1985a.) In contrast, DDT, a contaminant that accumulates in
fatty tissues, can biomagnify in many different types of food chains. Upper-trophic-level
species (such as predatory birds), therefore, are likely to be exposed to higher levels of DDT
through their prey than are lower-trophic-level species in the ecosystem.
3.5 SELECTION OF ASSESSMENT ENDPOINTS
As noted in the introduction to this guidance, an assessment endpoint is "an explicit
expression of the environmental value that is to be protected" (U.S. EPA, 1992a). In human
health risk assessment, only one species is evaluated, and cancer and noncancer effects are the
usual assessment endpoints. Ecological risk assessment, on the other hand, involves multiple
species that are likely to be exposed to differing degrees and to respond differently to the
same contaminant. Nonetheless, it is not practical or possible to directly evaluate risks to all
of the individual components of the ecosystem at a site. Instead, assessment endpoints focus
the risk assessment on particular components of the ecosystem that could be adversely
affected by contaminants from the site.
The selection of assessment endpoints includes discussion between the lead risk
assessor and the risk manager concerning management policy goals and ecological values.
The lead risk assessor and risk manager should seek input from the regional BTAG, PRPs,
and other stakeholders associated with a site when identifying assessment endpoints for a site.
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Stakeholder input at this stage will help ensure that the risk manager can readily defend the
assessment endpoints when making decisions for the site. ECO Update Volume 3, Number 1,
briefly summarizes the process of selecting assessment endpoints (U.S. EPA, 1995b).
Individual assessment endpoints usually encompass a group of species or populations
with some common characteristics, such as a specific exposure route or contaminant
sensitivity. Sometimes, individual assessment endpoints are limited to one species (e.g., a
species known to be particularly sensitive to a site contaminant). Assessment endpoints can
also encompass the typical structure and function of biological communities or ecosystems
associated with a site.
Assessment endpoints for the baseline ecological risk assessment must be selected
based on the ecosystems, communities, and/or species potentially present at the site. The
selection of assessment endpoints depends on:
(1) The contaminants present and their concentrations;
/
(2) Mechanisms of toxicity of the contaminants to different groups of organisms;
(3) Ecologically relevant receptor groups that are potentially sensitive or highly
exposed to the contaminant and attributes of their natural history; and
(4) Potentially complete exposure pathways.
Thus, the process of selecting assessment endpoints can be intertwined with other phases of
problem formulation.
The risk assessment team must think through the contaminant mechanism(s) of
ecotoxicity to determine what receptors will or could be at risk. This understanding must
include how the adverse effects of the contaminants might be expressed (e.g., eggshell
thinning in birds), as well as how the chemical and physical form of the contaminants
influence bioavailability and the type and magnitude of adverse response (e.g., inorganic
versus organic mercury).
The risk assessment team also should determine if the contaminants can adversely
affect organisms in direct contact with the contaminated media (e.g., direct exposure to water,
sediment, soil) or if the contaminants accumulate in food chains, resulting in adverse effects
in organisms that are not directly exposed or are minimally exposed to the original
contaminated media (indirect exposure). The team should decide if the risk assessment
should focus on toxicity resulting from direct or indirect exposures, or if both must be
evaluated.
Broad assessment endpoints (e.g., protecting aquatic communities) are generally of less
value in problem formulation than specific assessment endpoints (e.g., maintaining aquatic
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community composition and structure downstream of a site similar to that upstream of the
site). Specific assessment endpoints define the ecological value in sufficient detail to identify
the measures needed to answer specific questions or to test specific hypotheses. Example 3-3
provides three examples of assessment endpoint selection based on the hypothetical sites in
Appendix A.
The formal identification of assessment endpoints is part of the SMDP for this step.
Regardless of the level of effort to be expended on the subsequent phases of the risk
assessment, the assessment endpoints identified are critical elements in the design of the
ecological risk assessment and must be agreed upon as the focus of the risk assessment.
Once assessment endpoints have been selected, testable hypotheses and measurement
endpoints can be developed to determine whether or not a potential threat to the assessment
endpoints exists. Testable hypotheses and measurement endpoints cannot be developed
without agreement on the assessment endpoints among the risk manager, risk assessors, and
other involved professionals.
3.6 THE CONCEPTUAL MODEL AND RISK QUESTIONS
The site conceptual model establishes the complete exposure pathways that will be
evaluated in the ecological risk assessment and the relationship of the measurement endpoints
to the assessment endpoints. In the conceptual model, the possible exposure pathways are
depicted in an exposure pathway diagram and must be linked directly to the assessment
endpoints identified in Section 3.5. Developing the conceptual model and risk questions are
described in Sections 3.6.1 and 3.6:2, respectively. Selection of measurement endpoints,
completing the conceptual model, is described in Step 4.
3.6.1 Conceptual Model
Based on the information obtained from Steps 1 and 2, knowledge of the contaminants
present, the exposure pathway diagram, and the assessment endpoints, an integrated
conceptual model is developed (see Example 3-4). The conceptual model includes a
contaminant fate-and-transport diagram that traces the contaminants' movement from sources
through the ecosystem to receptors that include the assessment endpoints (see Example 3-5).
Contaminant exposure pathways that do not lead to a species or group of species associated
with the proposed assessment endpoint indicate that either:
(1) There is an incomplete exposure pathway to the receptor(s) associated with the
proposed assessment endpoint; or
(2) There are missing components or data necessary to demonstrate a complete
exposure pathway.
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EXAMPLE 3-3
Assessment Endpoint Selection-DDT, Copper, and PCB Sites
DDT Site
An assessment endpoint such as "protection of the ecosystem from the effects of DDT"
would give little direction to the risk assessment. However, "protection of piscivorous birds
from eggshell thinning due to DDT exposure" directs the risk assessment toward the food-chain
transfer of DDT that results in eggshell thinning in a specific group of birds. This assessment
endpoint provides the foundation for identifying appropriate measures of effect and exposure
and ultimately the design of the site investigation. It is not necessary that a specific species of
bird be identified on site. It is necessary that the exposure pathway exists and that the presence
of a piscivorous bird could be expected.
Copper Site
Copper can be acutely or chronically toxic to organisms in an aquatic community
through direct exposure of the organisms to copper in the water and sediments. Threats of
copper toxicity to higher-trophic-level organisms are unlikely to exceed threats to organisms at
the base of the food chain, because copper is an essential nutrient which is effectively regulated
by most organisms if the exposure is below immediately toxic levels. Aquatic plants
(particularly phytoplankton) and mollusks, however, are poor at regulating copper and might be
sensitive receptors or effective in transferring copper to the next trophic level. In addition, fish
fry can be very sensitive to copper in water. Based on these receptors and the potential for both
acute and chronic toxicity, an appropriate general assessment endpoint for the system could be
the maintenance of aquatic community composition. An operational definition of the
assessment endpoint for this site would be pond fish and invertebrate community composition
similar to that of other ponds of similar size and characteristics in the area.
PCB Site
The primary ecological threat of PCBs in ecosystems is not through direct exposure and
acute toxicity. Instead, PCBs bioaccumulate in food chains and can diminish reproductive
success in some vertebrate species. PCBs have been implicated as a cause of reduced
reproductive success of piscivorous birds (e.g., cormorants, terns) in the Great Lakes (Kubiak et
al., 1989; Fox et al., 1991) and of mink along several waterways (Aulerich and Ringer, 1977;
Foley et al., 1988). Therefore, reduced reproductive success in high-trophic-level species
exposed via their diet is a more appropriate assessment endpoint than either toxicity to
organisms via direct exposure to PCBs in water, sediments, or soils, or reproductive impairment
in lower-trophic-level species.
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EXAMPLE 3-4
Description of the Conceptual Model-DDT Site
One of the assessment endpoints selected for the DDT site (Appendix A) is the
protection of piscivorous birds. The site 'conceptual model includes the release of DDT from
the spill areas to the adjacent stream, followed by food chain accumulation of DDT from the
sediments and water through the lower trophic levels to forage fish in the stream. The forage
fish are the exposure point for piscivorous birds. Eggshell thinning was selected as the measure
of effect. During the literature review of the ecological effects of DDT, toxicity studies were
found that reported reduced reproductive success (i.e., number of young fledged) in birds that
experienced eggshell thinning of 20 percent or more (Anderson and Hickey, 1972; Dilwqrth et
al., 1972). Based on those data, the lead risk assessor and risk manager agreed that eggshell
thinning of 20 percent or more would be considered an adverse effect for piscivorous birds.
Chronic DDT exposure can also reduce some animals' ability to escape predation.
Thus, DDT can indirectly increase the mortality rate of these organisms by making them more
susceptible to predators (Cooke, 1971; Krebs et al., 1974). That effect of DDT oh prey also can
have an indirect consequence for the predators. If predators are more likely to capture the more
contaminated prey, the predators could be exposed to DDT at levels higher than represented in
the average prey population.
If case (1) is true, the proposed assessment endpoint should be reevaluated to determine if it
is an appropriate endpoint for the site. If case (2) is true, then additional field data could be
needed to evaluate contaminant fate and transport at the site. Failure to identify a complete
exposure pathway that does exist at the site can result in incorrect conclusions or in extra
time and effort being expended on a supplementary investigation.
As indicated in Section 3.5, appropriate assessment endpoints differ from site to site,
and can be at one or more levels of biological organization. At any particular site, the
appropriate assessment endpoints might involve local populations of a particular species,
community-level integrity, and/or habitat preservation. The site conceptual model must
encompass the level of biological organization appropriate for the assessment endpoints for
the site. The conceptual model can use assumptions that generally represent a group of
organisms or ecosystem components.
The intent of the conceptual model is not to describe a particular species or site
exactly as much as it is to be systematic, representative, and conservative where information
is lacking (with assumptions biased to be more likely to overestimate than to underestimate
risk). For example, it is not necessary or even recommended to develop new test protocols to
use species that exist at a site to test the toxicity of site media (See Step 4). Species used in
standardized laboratory toxicity tests (e.g., fathead minnows, Hyallela amphipods) usually are
adequate surrogates for species in their general taxa and habitat at the site.
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EXAMPLE 3-5
Conceptual Model Diagram-DDT Site
SECONDARY
RECEPTOR
(Fish)
u*
PRIMARY SOURCE
(Plant site)
SECONDARY
SOURCE
(Surface drainage)
TERTIARY SOURCE
(Stream sediments,
exposure point for fish and
macroinvertebrates)
ASSESSMENT
ENDPOINT
TERTIARY RECEPTOR
(Piscivorous bird)
PRIMARY RECEPTOR
(Benthic
macroinvertebrates,
exposure point for fish)
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3.6.2 Risk Questions
Ecological risk questions for the
baseline risk assessment at Superfund .sites
are basically questions about the
relationships among assessment endpoints
and their predicted responses when exposed
to contaminants. The risk questions should
be based on the assessment endpoints and
provide a basis for developing the study
design (Step 4) and for evaluating the
results of the site investigation in the
analysis phase (Step 6) and during risk
characterization (Step 7).
The most basic question applicable
to virtually all Superfund sites is whether
site-related contaminants are causing or have
the potential to cause adverse effects on the
assessment endpoint(s). To use the baseline
ecological risk assessment in the FS to
evaluate remedial alternatives, it is helpful if
the specific contaminant(s) responsible can
be identified. Thus refined, the question
becomes "does (or could) chemical X cause adverse effects on the assessment endpoint?" In
general, there are four lines of evidence that can be used to answer this question:
(1) Comparing estimated or measured exposure levels to chemical X with levels that
are known from the literature to be toxic to receptors associated with the
assessment endpoints;
(2) Comparing laboratory bioassays with media from the site and bioassays with media
from a reference site;
(3) Comparing in situ toxicity tests at the site with in situ toxicity tests in a reference
body of water; and
(4) Comparing observed effects in the receptors associated with the site with similar
receptors at a reference site.
These lines of evidence are considered further in Step 4, as measurement endpoints are
selected to complete the conceptual model and the site-specific study is designed.
HIGHLIGHT 3-3
Definitions:
Null and Test Hypotheses
Null hypothesis: Usually a hypothesis of
no differences between two populations
formulated for the express purpose of being
rejected.
Test (or alternative) hypothesis: An
operational statement of the investigator's
research hypothesis.
"When appropriate, formal hypothesis
testing is preferred to make explicit what
error rates are acceptable and what
magnitude of effect is considered
biologically important. However, it might
not be practical for many assessment
endpoints or be the only acceptable way to
state questions about those endpoints. See
Example 4-1 in the next chapter.
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3.7 SCIENTIFIC/MANAGEMENT DECISION POINT (SMDP)
At the conclusion of Step 3, there is .a SMDP. The SMDP consists of agreement on
four items: contaminants of concern, assessment endpoints, exposure pathways, and risk
questions. Those items can be summarized with the assistance of the diagram of the
conceptual model. Without agreement between the risk manager, risk assessors, and other
involved professionals on the conceptual model to this point, measurement endpoints cannot
be selected, and a site study cannot be developed effectively. Example 3-5 shows the
conceptual model for the DDT site example in Appendix A.
3.8 SUMMARY
By combining information on: (1) the potential contaminants present; (2) the
ecotoxicity of the contaminants; (3) environmental fate and transport; (4) the ecological
setting; and (5) complete exposure pathways, an evaluation is made of what aspects of the
ecosystem at the site could be at risk and what the adverse ecological response could be.
"Critical exposure pathways" are based on: (1) exposure pathways to sensitive species'
populations or communities; and (2) exposure levels associated with predominant fate and
transport mechanisms at a site.
Based on that information, the risk assessors and risk manager agree on assessment
endpoints and specific questions or testable hypotheses that, together with the rest of the
conceptual model, form the basis for the site investigation. At this stage, site-specific
information on exposure pathways and/or the presence of specific species is likely to be
incomplete. By using the conceptual model developed thus far, measurement endpoints can
be selected, and a plan for filling information gaps can be developed and written into the
ecological WP and SAP as described in Step 4.
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STEP 4: STUDY DESIGN AND DATA QUALITY
OBJECTIVE PROCESS
OVERVIEW
The site conceptual model begun in Step 3, which includes assessment
endpoints, exposure pathways, and risk questions or hypotheses, is completed in Step 4
with the development of measurement endpoints. The conceptual model then is used
to develop the study design and data quality objectives. The products of Step 4 are the
ecological risk assessment WP and SAP, which describe the details of the site
investigation as well as the data analysis methods and data quality objectives (DQOs).
As part of the DQO process, the SAP specifies acceptable levels of decision errors that
will be used as the basis for establishing the quantity and quality of data needed to
support ecological risk management decisions.
The lead risk assessor and the risk manager should agree that the WP and SAP
describe a study that will provide the risk manager with the information needed to
fulfill the requirements of the baseline risk assessment and to incorporate ecological
considerations into the site remedial process. Once this step is completed, most of the
professional judgment needed for the ecological risk assessment will have been
incorporated into the design and details of the WP and SAP. This does not limit the
need for qualified professionals in the implementation of the investigation, data
acquisition, or data interpretation. However, there should be no fundamental changes
in goals or approach to the ecological risk assessment once the WP and SAP are
finalized.
It is important to coordinate this step with the WP and SAP for the site
investigation, which is used to document the nature and extent of contamination and to
evaluate human health risks.
Step 4 of the ecological risk assessment establishes the measurement endpoints
(Section 4.1), completing the conceptual model begun in Step 3. Step 4 also establishes the
study design (Section 4.2) and data quality objectives based on statistical considerations
(Section 4.3) for the site assessment that will accompany site-specific studies for the remedial
investigation! The site conceptual model is used to identify which points or assumptions in
the risk assessment include the greatest degree of conservatism or uncertainty. The field
sampling then can be designed to address the risk model parameters that have important
effects on the risk estimates (e.g., bioavailability and toxicity of contaminants in the field,
contaminant concentrations at exposure points).
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The products of Step 4 are the WP and SAP for the ecological component of the field
investigations (Section 4.4). Involvement of the BTAG in the preparation, review, and
approval of WPs and SAPs can help ensure that the ecological risk assessment is well
focused, performed efficiently, and technically correct.
The WP and SAP should specify the site conceptual model developed in Step 3, and
the measurement endpoints developed in the beginning of Step 4. The WP describes:
• Assessment endpoints;
• Exposure pathways;
• Questions and testable hypotheses;
• Measurement endpoints and their relation to assessment endpoints; and
• Uncertainties and assumptions.
The SAP should describe:
• Data needs;
• Scientifically valid and sufficient study design and data analysis procedures;
• Study methodology and protocols, including sampling techniques;
• Data reduction and interpretation techniques, including statistical analyses; and
• Quality assurance procedures and quality control techniques.
The SAP must include the data reduction and interpretation techniques, because it is necessary
to known how the data will be interpreted to specify the number of samples needed.
Prior to formal agreement on the WP and SAP, the proposed field sampling plan is
verified in Step 5.
4.1 ESTABLISHING MEASUREMENT ENDPOINTS
As indicated in the Introduction, a measurement endpoint is defined as "a measurable
ecological characteristic that is related to the valued characteristic chosen as the assessment
endpoint" and is a measure of biological effects (e.g., mortality, reproduction, growth) (U.S.
EPA, 1992a; although this definition may change—see U.S. EPA 1996a). Measurement
endpoints are frequently numerical expressions of observations (e.g., toxicity test results,
community diversity measures) that can be compared statistically to a control or reference site
to detect adverse responses to a site contaminant. As used in this guidance, measurement
endpoints can include measures of exposure (e.g., contaminant concentrations in water) as
well as measures of effect. The relationship between measurement and assessment endpoints
must be clearly described within the conceptual model and must be based on scientific
evidence. This is critical because the assessment and measurement endpoints usually are
different endpoints (see the Introduction and Highlight 4-1).
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Typically, the number of
measurement endpoints that are potentially
appropriate for any given assessment
endpoint and circumstance is limited. The
most appropriate measurement endpoints for
an assessment endpoint depend on several
considerations, a primary one being how
many and which lines of evidence are
needed to support risk-management
decisions at the site (see Section 3.6.2).
Given the potential ramifications of site
actions, the site risk manager might want to
use more than one line of evidence to
identify site-specific thresholds for effects.
The risk manager and risk assessors must
consider the utility of each type of data
given the cost of collecting those data and
the likely sensitivity of the risk estimates to
the data.
HIGHLIGHT 4-1
Importance of Distinguishing
Measurement from Assessment
Endpoints
If a measurement endpoint is
mistaken for an assessment endpoint, the
misperceptionvcan arise that Superfund is
basing a remediation on an arbitrary or
esoteric justification. For example,
protection of a few invertebrate and algal
species could be mistaken as the basis for a
remedial decision, when the actual basis for
the decision is the protection of the aquatic
community as a whole (including higher-
trophic-level game fish that depend on lower
trophic levels in the community), as
indicated by a few sensitive invertebrate and
algal species.
There are some situations in which it
might only be necessary or possible to compare estimated or measured contaminant exposure
levels at a site to ecotoxicity values derived from the literature. For example, for
contaminants in surface waters for which there are state water-quality standards, exceedance
of the standards indicates that remediation to reduce contaminant concentrations in surface
waters to below these levels could be needed whether impacts are occurring or not. For
assessment endpoints for which impacts are difficult to demonstrate in the field (e.g., because
of high natural variability), and toxicity tests are not possible (e.g., food-chain accumulation is
involved), comparing environmental concentrations with a well-supported ecotoxicity value
might have to suffice.
A bioassay using contaminated media from the site can suffice if the risk manager and
risk assessor agree that laboratory tests with surrogate species will be taken as indicative of
likely effects on the assessment endpoint. For sites with complex mixtures of contaminants
without robust ecotoxicity values and high natural variability in potential measures for the
assessment endpoint, either laboratory or in situ toxicity testing might be the best technique
for evaluating risks to the assessment endpoint. For inorganic substances in soils or
sediments, bioassays often are needed to determine the degree to which a contaminant is
bioavailable at a particular site. Laboratory toxicity tests can indicate the potential for
adverse impacts in the field, while in situ toxicity testing with resident organisms can provide
evidence of actual impacts occurring in the field.
Sometimes more than one line of evidence is needed to reasonably demonstrate that
contaminants from a site are likely to cause adverse effects on the assessment endpoint. For
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example, total recoverable copper in a surface water body to which a water quality standard
did not apply could exceed aquatic ecotoxicity values, but not cause adverse effects because
the copper is only partially bioavailable or because the ecotoxicity value is too conservative
for the particular ecosystem. Additional evidence from bioassays or community surveys
could help resolve whether the copper is actually causing adverse effects (See Example 4-1).
Alternatively, if stream community surveys indicate impairment of community structure
downstream of a site, comparing contaminant concentrations with aquatic toxicity values can
help identify which contaminants are most likely to be causing the effect. When some lines
of evidence conflict with others, professional judgment is needed to determine which data
should be considered more reliable or relevant to the questions.
EXAMPLE 4-1
Lines of Evidence-Copper Site
Primary question: Are ambient copper levels in sediments causing adverse effects in
benthic organisms in the pond?
Possible lines of evidence phrased as test hypotheses:
x
(1) Mortality in early life stages of benthic aquatic insects in contact with
sediments from the site significantly exceeds mortality in the same kinds
of organisms in contact with sediments from a reference site (e.g.,
(2) Mortality in in situ toxicity tests in sediments at the pond significantly
exceeds mortality in in situ toxicity tests in sediments at a reference pond
(e.g., p< 0.1).
(3) There are significantly fewer numbers of benthic aquatic insect species
present per m2 of sediment at the pond near the seep than at the opposite
side of the pond (e.g., p < 0.1).
Statistical and biological significance: Differences in the incidence of adverse
effects between groups of organisms exposed to contaminants from the site and groups
not exposed might be statistically significant, but not biologically important, depending
on the endpoint and the power of the statistical test. Natural systems can sustain some
level of perturbation without changing in structure or function. The risk assessor needs
to evaluate what level of effect will be considered biologically important. Given the
limited power of small sample sizes to detect an effect, the risk assessor might decide
that any difference that is statistically detectable at a p level of 0.1 or lessjs important
biologically.
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Once there is agreement on which lines of evidence are required to answer questions
concerning the assessment endpoint, the measurement endpoints by which the questions or
test hypotheses will be examined can be selected.
Each measurement endpoint should represent the same exposure pathway and toxic
mechanism of action as the assessment endpoint it represents; otherwise, irrelevant exposure
pathways or toxic mechanisms might be evaluated. For example, if a contaminant primarily
causes damage to vertebrate kidneys, the use of daphnids (which do not have kidneys) would
be inappropriate. ' . ,
Potential measurement endpoints in toxicity tests or in field studies should be
evaluated according to how well they can answer questions about the assessment endpoint or
support or refute the hypotheses developed for the conceptual model. Statistical
considerations, including sample size and statistical power described in Section 4.3, also must
be considered in selecting the measurement endpoints. The following subsections describe
additional considerations for selecting measurement endpoints, including
species/community/habitat (Section 4.1.1), relationship to the contaminant(s) of concern
(Section 4.1.2), and mechanisms of ecotoxicity (Section 4.1.3).
i
4.1.1 Species/Community/Habitat Considerations
The function of a measurement endpoint is to represent an assessment endpoint for the
site. The measurement endpoint must allow clear inferences about potential changes in the
assessment endpoint. Whenever assessment and measurement endpoints are not the same
(which usually is the case), measurement endpoints should be selected to be inclusive of risks
to all of the species, populations, or groups included in the assessment endpoint that are not
directly measured. In other words, the measurement endpoint should be representative of the
assessment endpoint for the site and not lead to an underestimate of risk to the assessment
endpoint. Example 4-2 illustrates this point for the DDT site in Appendix A.
In selecting a measurement endpoint, the species and life stage, population, or
community chosen should be the one(s) most susceptible to the contaminant for the
assessment endpoint in question. For species and populations, this selection is based on a
review of the species: (1) life history; (2) habitat utilization; (3) behavioral characteristics;
and (4) physiological parameters. Selection of measurement endpoints also should be based
on which routes of exposure are likely. For communities, careful evaluation of the
contaminant fate and transport in the environment is essential.
4.1.2 Relationship of the Measurement Endpoints to the Contaminant of
Concern
Additional criteria to consider when selecting measurement endpoints are inherent
properties (such as the physiology or behavioral characteristics of the species) or life history
parameters that make a species useful in evaluating the effects of site-specific contaminants.
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For example, Chironomus tentans (a
species of midge that is used as a
standard sediment toxicity testing
species in the larval stage) is
considered more tolerant of metals
contamination than is C. riparius, a
similar species (Klemm et al., 1990;
Nebeker et al., 1984; Pascoe et al.,
1989). To assess the effects of
exposure of benthic communities to
metal-contaminated sediment, C.
riparius might be the better species to
use as a test organism for many aquatic
systems to ensure that risks are not
underestimated. In general, the most
sensitive of the measurement endpoints
appropriate for inferring risks to the
assessment endpoint should be used. If
all else is equal, however, species that are commonly used in the laboratory are preferred over
non-standard laboratory species to improve test precision.
Some species have been identified as being particularly sensitive to certain
contaminants. For example, numerous studies have demonstrated that mink are among the
most sensitive of the tested mammalian species to the toxic effects of PCBs (U.S. EPA,
1995a). Species that rely on quick reactions or behavioral responses to avoid predators can
be particularly sensitive to contaminants affecting the central nervous system, such as
mercury. Thus, the sensitivity of the measurement endpoint relative to the assessment
endpoint should be considered for each contaminant of concern.
HIGHLIGHT 4-2
Terminology and Definitions
In the field of ecotoxicology, there
historically have been multiple definitions for
some terms, including definitions for direct
effects, indirect effects, acute effects, chronic
effects, acute tests, and chronic tests. This
multiplicity of definitions has resulted in
misunderstandings and inaccurate communication
of study designs. Definitions of these and other
terms, as they are used in this document, are
provided in the glossary. When consulting other
reference materials, the user should evaluate how
the authors defined terms.
EXAMPLE 4-2
Selecting Measurement Endpoints-DDT Site
As described in Example 3-1, one of the assessment endpoints selected for the DDT site
is the protection of piscivorous birds from egg-shell thinning due to DDT exposure. The belted
kingfisher was selected as a piscivorous bird with the smallest home range that could utilize the
area of the site, thereby maximizing the calculated dose to a receptor. In this illustration, the
kingfishers are used as the most highly exposed of the piscivorous birds potentially present.
Thus, one can conclude that, if the risk assessment shows no threat of eggshell thinning to the
kingfisher, there should be minimal or no threat to other piscivorous birds that might utilize the
site. Thus, eggshell thinning in belted kingfishers is an appropriate measurement endpoint for
this site.
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4.1.3 Mechanisms of Ecoxicity
A contaminant can exert adverse ecological effects in many ways. First, a
contaminant might affect an organism after exposure for a short period of time (acute) or after
exposure over an extended period of time (chronic). Second, the effect of a contaminant
could be lethal (killing the organism) or sublethal (causing adverse effects other than death,
such as reduced growth, behavioral changes, etc.). Sublethal effects can reduce an organism's
lifespan or reproductive success. For example, if a contaminant reduces the reaction speed of
a prey species, the prey can become more susceptible to predation. Third, a contaminant
might act directly or indirectly on an organism. Direct effects include lethal or sublethal
effects of the chemical on the organism. Indirect effects occur when the contaminant
damages the food, habitat, predator-prey relationships, or competition of the organism in its
community.
Mechanisms of ecotoxicity and exposure pathways have already been considered
during problem formulation and identification of the assessment endpoints. However, toxicity
issues are revisited when selecting appropriate measurement endpoints to ensure that they
measure the assessment endpoint's toxic response of concern.
4.2 STUDY DESIGN
In Section 4.1, one or more lines of evidence that could be used to answer questions
or to test hypotheses concerning the assessment endpoint(s) were identified. This section
provides recommendations on how to design a field study for: bioaccumulation and field ^
tissue residue studies (Section 4.2.1); population/community evaluations (Section 4.2.2); and
toxicity testing (Section 4.2.3). A thorough understanding of the strengths and limitations of
these types of field studies is necessary to properly design any investigation.
Typically, no one line of evidence can stand on its own. Analytic chemistry on co-
located samples and other lines of evidence are needed to support a conclusion. When
population/community evaluations are coupled with toxicity testing and media chemistry, the
procedure often is referred to as a triad approach (Chapman et al., 1992; Long and Chapman,
1985). This method has proven effective in defining the area affected by contaminants in
sediments of several large bays and estuaries.
The development of exposure-response relationships is critical for evaluating risk
'management options; thus, for all three types of studies, sampling is applied to a
contamination gradient when possible as well as compared to reference data. Reference data
are baseline values or characteristics that should represent the site in the absence of
contaminants released from the site. Reference data might be data collected from the site
before contamination occurred or new data collected from a reference site. The reference site
can be the least impacted (or unimpacted) area of the Superfund site or a. nearby site that is
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ecologically similar, but not affected by the site's contaminants. For additional information
on selecting and using reference information in Superfund ecological risk assessments, see
ECO Update Volume 2, Number 1 (U.S. EPA, 1994e).
The following subsections present a starting point for selecting an appropriate study
design for the different types of biological sampling that might apply to the site investigation.
4.2.1 Bioaccumulation and Field Tissue Residue Studies
Bioaccumulation and field tissue residue studies typically are conducted at sites where
contaminants are likely to accumulate in food chains; The studies help to evaluate
contaminant exposure levels associated with measures of effect for assessment endpoint
species.
The degree to which a contaminant is transferred through a food chain can be
evaluated in several ways. The most common type of study reported in the literature is a
contaminant bioaccumulation (uptake) study. As indicated in Section 2.2.1, the most
conservative BAF values identified in the literature generally are used to estimate
bioaccumulation in Step 2 of the screening-level risk assessment. Where the potential for
overestimating bioaccumulation by using conservative literature values to represent the site is
substantial, additional evaluation of the literature for values more likely to apply to the site or
a site-specific tissue residue study might be advisable.
A tissue residue study generally is conducted on organisms that are in the exposure
pathway (i.e., food chain) associated with the assessment endpoint. Data seldom are available
to link tissue residue levels in the sampled organisms to adverse effects in those organisms.
Literature toxiciry studies usually associate effects with an administered dose (or data that can
be converted to an administered dose), not a tissue residue level. Thus, the purpose of a field
tissue residue study usually is to measure contaminant concentrations in foods consumed by
the species associated with the assessment endpoint. This measurement minimizes the
uncertainty associated with estimating a dose (or intake) to that species, particularly in
situations in which several media and trophic levels are in the exposure pathway.
The concentration of a contaminant in the primary prey/food also should be linked to
an exposure concentration from a contaminated medium (e.g., soil, sediment, water), because
it is the medium, not the food chain, that will be remediated. Thus, contaminant
concentrations must be measured in environmental media at the same locations at which the
organisms are collected along contaminant gradients and at reference locations. Co-located
samples of the contaminated medium and organisms are needed to establish a correlation
between the tissue residue levels and contamination levels in the medium under evaluation;
these studies are most effective if conducted over a gradient of contaminant concentrations.
In addition, tissue residues from sessile organisms (e.g., rooted plants, clams) are easier to
attribute to specific contaminated areas than are tissue residues from mobile organisms (e.g..
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large fish). Example 4-3 illustrates these concepts using the DDT site example in
Appendix A.
EXAMPLE 4-3
Tissue Residue Studies-DDT Site
In the DDT site example, a forage fish (e.g., creek chub) will be collected at
several locations with known DDT concentrations in sediments. The forage fish will be
analyzed for body burdens of DDT, and the relationship between the DDT levels in the
sediments and the levels in the forage fish will be established. The forage fish DDT
concentrations can be used to evaluate the DDT threat to piscivorous birds feeding on
the forage fish at each location. Using the DDT concentrations measured in fish that
correspond to a LOAEL and NOAEL for adverse effects in birds and the relationship
between the DDT levels in the sediments and in the forage fish, the corresponding
sediment contamination levels can be estimated. Those sediment DDT concentrations
can then be used to estimate a cleanup level that would reduce threats of eggshell
thinning to piscivorous birds.
Although it might seem obvious, it is important to confirm that the organisms
examined for tissue residue levels are in the exposure pathways of concern established by the
conceptual model. Food items targeted for collection should be those that are likely to
constitute a large portion of the diet of the species of concern (e.g., new growth on maple
trees, rather than cattails, as a food source for deer) and/or represent pathways of maximum
exposure. If not, erroneous conclusions or study delays and added costs can result. Because
specific organisms often can only be captured in one season, the timing of the study can be
critical, and failure to plan accordingly can result in serious site management difficulties.
There are numerous factors that must be considered when selecting a species in which
to measure contaminant residue levels. Several investigators have discussed the "ideal"
characteristics of the species to be collected and analyzed. The recommendations of Phillips
(1977, 1978) include that the species selected should be:
(I) Able to accumulate the chemical of concern without being adversely affected
by the levels encountered at the site;
(2) Sedentary (small home range) in order to be representative of the area of
collection;
(3) Abundant in the study area; and
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(4) Of reasonable size to give adequate tissue for analysis (e.g., 10 grams for
organic analysis and 0.5 gram for metal analysis for many laboratories (Roy F.
Weston, Inc., 1994)).
Additional considerations for some situations would be that the species is:
(5) Sufficiently long-lived to allow for sampling more than one age class; and
(6) Easy to sample and hardy enough to survive in the laboratory (allowing for the
organisms to eliminate contaminants from their gastrointestinal tract prior to
analysis, if desired, and allowing for laboratory studies on the uptake of the
contaminant).
It is usually not possible or necessary to find an organism that fulfills all of the above
requirements. The selection of an organism for tissue analysis should balance these
characteristics with the hypotheses being tested, knowledge of the contaminants' fate and
transport, and the practicality of using the particular species. In the following sections,
several of the factors mentioned above are described in greater detail.
Ability to accumulate the contaminant. The objectives of a tissue residue study
are (1) to measure bioavailability directly; (2) to provide site-specific estimates of exposure to
higher-trophic-level organisms; and (3) to relate tissue residue levels to concentrations in
environmental media (e.g., in soil, sediment, or water). Sometimes these studies also can be
used to link tissue residue levels with observed effects in the organisms sampled. However,
in a "pure" accumulation study, the species selected for collection and tissue analysis should
be ones that can accumulate a contaminant(s) without being adversely affected by the levels
encountered in the environment. While it is difficult to evaluate whether or not a population
in the field is affected by accumulation of a contaminant, it is important,to try. Exposure that
results in adverse responses might alter the animal's feeding rates or efficiency, diet, degree
of activity, or metabolic rate, and thereby influence the animal's daily intake or accumulation
of the contaminant and the estimated BAF. For example, if the rate of bioaccumulation of a
contaminant in an organism decreases with increasing environmental concentrations (e.g., its
toxic effects reduce food consumption rates), using a BAF determined at low environmental
concentrations to estimate bioaccumulation at high environmental concentrations would
overestimate risk. Conversely, if bioaccumulation increased with increasing environmental
concentrations (e.g., its toxic effects impair the organisms' ability to excrete the contaminant),
using a BAF determined at low environmental concentrations would underestimate risks at
higher environmental concentrations.
Consideration of the physiology and biochemistry of the species selected for residue
analysis also is important. Some species can metabolize certain organic contaminant(s) (e.g.,
fish can metabolize PAHs). If several different types of prey are consumed by a species of
concern, it would be more appropriate to analyze prey species that do not metabolize the
contaminant.
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Home range. When selecting species for residue analyses, one should be confident
that the contaminant levels found in the organism depend on the contaminant levels in the
environmental media under evaluation. Otherwise, valid conclusions cannot be drawn about
ecological risks posed by contaminants at the site. The home range, particularly the foraging
areas within the home range, and movement patterns of a species are important in making this
determination. Organisms do not utilize the environment uniformly. For species that have
large home ranges or are migratory, it can be difficult to evaluate potential exposure to
contaminants at the site. Attribution of contaminant levels in an organism to contaminant
levels in the surrounding environment is easiest for animals with small home and foraging
ranges and limited movement patterns. Examples of organisms with small home ranges
include young-of-the-year fish, burrowing Crustacea (such as fiddler crabs or some crayfish),
and small mammals.
Species also should be selected for residue analysis to maximize the overlap between
the area of contamination and the species' home range or feeding range. This provides a
conservative evaluation of potential exposure levels. The possibility that a species' preferred
foraging areas within a home range overlap the areas of maximum contamination also should
.be considered.
Population size. A species selected for tissue residue analysis should be sufficiently
abundant at the site that adequate numbers (and sizes) of individuals can be collected to
support the tissue mass requirements for chemical analysis and to achieve the sample size
needed for statistical comparisons. The organisms actually collected should be not only of
the same species, but also of similar age or size to reduce data variability when BAFs are
being evaluated. The practicality of using a particular species is evaluated in Step 5.
Size/composites. When selecting species in which to measure tissue residue levels,
it is best to have individual animals large enough for chemical analysis, without having to
pool (combine) individuals prior to chemical analysis. However, composite samples will be
needed if individuals from the species selected cannot yield sufficient tissue for the required
analytical methods. Linking contaminant levels in organisms to concentrations in
environmental media is easier if composites are made up of members of the same species,
sex, size, and age, and therefore exhibit similar accumulation characteristics. When deciding
whether or not to pool samples, it is important to consider what impact the loss of
information on variability of contaminant levels along these dimensions will have on data
interpretation. The size, age, and sex of the species collected should be representative of the
range of prey consumed by the species of concern.
Summary. Although it can be difficult to meet all of the suggested criteria for
selecting a species for tissue residue studies, an attempt should be made to meet as many
criteria as possible. No formula is available for ranking the factors in order of importance
within a particular site investigation because the ranking depends on the study objectives.
However, a key criterion is that the organism be sedentary or have a limited home range. It
is difficult to connect site contamination to organisms that migrate over great distances or that
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have extremely large home ranges. Further information on factors that can influence
bioaccumulation is available from the literature (e.g., Phillips. 1977, 1978; U.S. EPA, 1995d).
4.2.2 Population/Community Evaluations
Population/community evaluations, or biological field surveys, are potentially useful
for both contaminants that are toxic to organisms through direct exposure to the contaminated
medium and contaminants that bioaccumulate in food chains. In either case, careful
consideration must be given to the mechanism of contaminant effects. Since
population/community evaluations are "impact" evaluations, they typically are not predictive.
The release of the contaminant must already have occurred and exerted an effect in order for
the population/community evaluation to be an effective tool for a risk assessment.
Population and community surveys evaluate the current status of an ecosystem, often
using several measures of population or community structure (e.g., standing biomass, species
richness) or function (e.g., feeding group analysis). The most commonly used measures
include number of species and abundance of organisms in an ecosystem, although some
species are difficult to evaluate. It is difficult to detect changes in top predator populations
affected by bioaccumulation of substances in their food chain due to the mobility of top
predators. Some species, most notably insects, can develop a tolerance to contaminants
(particularly pesticides); in these cases, a population/community survey would be ineffective
for evaluating existing impacts. While population/community evaluations can be useful, the
risk assessors should consider the level of effort required as well as the difficulty in
accounting for natural variability.
A variety of population/community evaluations have been used at Superfund sites.
Benthic macroinvertebrate surveys are the most commonly conducted population/community
evaluations. There are methods manuals (e.g., U.S. EPA 1989c, 1990a) and publications that
describe the technical procedures for conducting these studies. In certain instances, fish
community evaluations have proven useful at Superfund sites. However, these investigations
typically are more labor-intensive and costly than a comparable macroinvertebrate study. In
addition, fish generally are not sensitive measures of the effects of sediment contamination,
because they usually are more mobile than benthic macroinvenebrates. Terrestrial plant
community evaluations have been used to a limited extent at Superfund sites. For those
surveys, it is important to include information about historical land use and physical habitat
disruption in the uncertainty analysis.
Additional information on designing field studies and on field study methods can be
found in ECO Update Volume 2. Number 3 (U.S. EPA, 1994d).
Although population- and community-level studies can be valuable, several factors can
confound the interpretation of the results. For example, many fish and small mammal
populations normally cycle in relation to population density, food availability, and other
factors. Vole populations have been known to reach thousands of individuals per acre and
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then to decline to as low as tens of individuals per acre the following years without an
identifiable external strcssor (Geller, 1979). It is important that the "noise of the system" be
evaluated so that the impacts attributed to chemical contamination at the site are not actually
the result of different, "natural" factors. Populations located relatively close to each other can
be affected independently: one might undergo a crash, while another is peaking. Physical
characteristics of a site can isolate populations so that one population level is not a good
indicator of another; for example, a paved highway can be as effective a barrier as a river,
and populations on either side can fluctuate independently. Failure to evaluate such issues
can result in erroneous conclusions. The level of effort required to resolve some of these
issues can make population/community evaluations impractical in some circumstances.
4.2.3 Toxicity Testing
The bioavailability and toxicity of site contaminants can be tested directly with
toxicity tests. As with other methods, it is critical that the media tested are in exposure
pathways relevant to the assessment endpoint. If the site conceptual model involves exposure
of benthic invertebrates to contaminated sediments, then a solid-phase toxicity test using
contaminated sediments (as opposed to a water-column exposure test) and an infaunal species
would be appropriate. As indicated earlier, the species tested and the responses measured
must be compatible with the mechanism of toxicity. Some common site contaminants are not
toxic to most organisms at the same environmental concentrations that threaten top predators
because the contaminant biomagnifies in food chains (e.g., PCBs); toxicity tests using
contaminated media from the site would not be appropriate for evaluating this type of
ecological threat.
There are numerous U.S. EPA methods manuals and ASTM guides and procedures for
conducting toxicity tests (see references in the Bibliography). While documented methods
exist for a wide variety of toxicity tests, particularly laboratory tests, the risk assessor must
evaluate what a particular toxicity test measures and, just as importantly, what it does not
measure. Questions to consider when selecting an appropriate toxicity test include:
(1) What is the mechanism of toxicity of the contaminant(s)?
(2) What contaminated media are being evaluated (water, soil, sediment)?
(3) What toxicity test species are available to test the media being evaluated?
(4) What life stage of the species should be tested?
(5) What should the duration of the toxicity test be?
(6) Should the test organisms be fed during the test?
(7) What endpoints should be measured?
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There are a limited number of toxicity tests that are readily available for testing
environmental media. Many of the aquatic toxicity tests were developed for the regulation of
aqueous discharges to surface waters. These tests are useful, but one must consider the
original purpose of the test
New toxicity tests are being developed continually and can be of value in designing a
Superfund site ecological risk assessment. However, when non-standard tests are used,
complete documentation of the specific test procedures is necessary to support use of the data.
In situ toxicity tests involve placing organisms in locations that might be affected by
site contaminants and in reference locations. Non-native species should not be used, because
of the risk of their release into the environment in which they could adversely affect (e.g.,
prey on or outcompete) resident species. In situ tests might provide more realistic evidence
of existing adverse effects than laboratory toxicity tests; however, the investigator has little
control over many environmental parameters and the experimental organisms can be lost to
adverse weather or other events (e.g., human interference) at the site or reference location.
For additional information on using toxicity tests in ecological risk assessments, see
ECO Update Volume 2, Numbers 1 and 2 (U.S. EPA, 1994b,c).
4.3 DATA QUALITY OBJECTIVES AND STATISTICAL CONSIDERATIONS
The SAP indicates the number and location of samples to be taken, the number of
replicates for each sampling location, and the method for determining sampling locations. In
specifying those parameters, the investigator needs to consider, among other things, the DQOs
and statistical methods that will be used to analyze the data.
4.3.1 Data Quality Objectives
The DQO process represents a series of planning steps that can be employed
throughout the development of the WP and SAP to ensure that the type, quantity, and quality
of environmental data to be collected during the ecological investigation are adequate to
support the intended application. Problem formulation in Steps 3 and 4 is essentially the
DQO process. By employing problem formulation and the DQO process, the investigator is
able to define data requirements and error levels that are acceptable for the investigation prior
to the collection of data. This approach helps ensure that results are appropriate and
defensible for decision making. The specific goals of the general DQO process are to:
• Clarify the study objective and define the most appropriate types of data to
collect;
• Determine the most appropriate field conditions under which to collect the data;
and
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• Specify acceptable levels of decision errors that will be used as the basis for
establishing the quantity and quality of data needed to support risk management
decisions.
As the discussion of Steps 3 and 4 indicates, those goals are subsumed in the problem
formulation phase of an ecological risk assessment. Several U.S. EPA publications provide
detailed descriptions of the DQO process (U.S. EPA, 1993c,d,f, 1994f). Because many of the
steps of the DQO process are already covered during problem formulation, the DQO process
should be reviewed by the investigator and applied as needed.
4.3.2 Statistical Considerations
Sampling locations can be selected "randomly" to characterize an area or non-
randomly, as along a contaminant concentration gradient. The way in which sampling
locations are selected determines which statistical tests, if any, are appropriate for evaluating
test hypotheses.
If a toxicity test is to be used to identify contaminant concentrations in the
environment associated with a threshold for adverse effects, the statistical power of the test is
important. The threshold for effects is assumed to be between the NOAEL and LOAEL of a
toxicity test (see Section 7.3.1). For toxicity tests that use a small number of test and control
organisms or for which the toxic response is highly variable, the increase in response rate of
the test animals compared with controls often must be relatively high (e.g., 30 to 50 percent
increase) for the response to be considered a LOAEL (i.e., statistically increased level of an
adverse response compared with control levels). If a NOAEL-to-LOAEL range that might
represent a 20 to 50 percent increase'in adverse effect is unacceptable (e.g., a population is
unlikely to sustain itself with an additional 40 percent mortality), then the power of the study
design must be increased, usually by increasing sample size, but sometimes by taking full
advantage of all available information to improve the power of the design (e.g., stratified
sampling, special tests for trends, etc.). A limitation on the use of toxicity values from the
literature is that often, the investigator does not discuss the statistical power of the study
design, and hence does not indicate the minimum statistically detectable effect level.
Appendix D describes additional statistical considerations, including a description of Type I
and Type II error, statistical power, statistical models, and power efficiency.
In evaluating the results of statistical analyses, one should remember that a statistically
significant difference relative to a control or reference population does not necessarily imply a
biologically important or ecologically significant difference (see Example 4-1).
4.4 CONTENTS OF WORK PLAN AND SAMPLING AND ANALYSIS PLAN
The WP and SAP for the ecological investigation should be developed as pan of the
initial RI sampling event if possible. If not, the WP and SAP can be developed as an
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additional phase of the site investigation. In either case, the format of the WP and SAP
should be similar to that described by U.S. EPA (1988a, 1989b). Accordingly, those
documents should be consulted when developing the ecological investigation WP and SAP.
The WP and SAP are typically written as separate documents. In that case, the WP
can be submitted for the risk manager's review so that any concerns with the approach can be
resolved prior to the development of the SAP. For some smaller sites, it might be more
practical to combine the two documents, in which case, the investigators should discuss the
overall objectives and approach with the risk manager to ensure that all parties agree.
The WP and SAP are briefly described in Sections 4.4.1 and 4.4.2, respectively. A
plan for testing the SAP before the site WP and SAP are signed and the investigation begins
is described in Section 4.4.3.
4.4.1 Work Plan
The purpose of the WP is to document the decisions and evaluations made during
problem formulation and to identify additional investigative tasks needed to complete the
evaluation of risks to ecological resources. As presented in U.S. EPA (1988a), the WP
generally includes the following: '
• A general overview and background of the site including the site's physical
setting, ecology, and previous uses;
• A summary and analysis of previous site investigations and conclusions;
• A site conceptual model, including an identification of the potential exposure
pathways selected for analysis, the assessment endpoints and questions or
testable hypotheses, and the measurement endpoints selected for analysis;
• The identification of additional site investigations needed to conduct the
ecological risk assessment; and
• A description of assumptions used and the major sources of uncertainty in the
site conceptual model and existing information.
The general scope of the additional sampling activities also is presented in the WP. A
detailed description of the additional sampling activities is presented in the SAP along with an
anticipated schedule of the site activities.
4.4.2 Sampling and Analysis Plan
The SAP typically consists of two components: a field sampling plan (FSP) and a
quality assurance project plan (QAPP). The FSP provides guidance for all field work by
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providing a detailed description of the sampling and data-gathering procedures to be used for
the project. The QAPP provides a description of the steps required to achieve the objectives
dictated by the intended use of the data.
Field sampling plan. The FSP provides a detailed description of the samples
needed to meet the objectives and scope of the investigation outlined in the WP. The FSP for
the ecological assessment should be detailed enough that a sampling team unfamiliar with the
site would be able to gather all the samples and/or required field data based on the guidelines
presented in the document. The FSP for the ecological investigation should include a
description of the following elements:
• Sampling type and objectives;
• Sampling location, timing, and frequency;
• Sample designation;
• Sampling equipment and procedures; and
• Sample handling and analysis.
A detailed description of those elements for chemical analyses is provided in Appendix B of
U.S. EPA (1988a). Similar specifications should be developed for the biological sampling.
Quality assurance project plan. The objective of the QAPP is to provide a
description of the policy, organization, functional activities, and quality control protocols
necessary for achieving the study objectives. Highlight 4-3 presents the elements typically
contained in a QAPP.
U.S. EPA has prepared guidance on
the contents of a QAPP (U.S. EPA, 1987a,
1988a, 1989a). Formal quality assurance
and quality control (QA/QC) procedures
exist for some types of ecological
assessments, for example, for laboratory
toxicity tests on aquatic species. For
standardized laboratory tests, there are
formal QA/QC procedures that specify (1)
sampling and handling of hazardous wastes;
(2) sources and culturing of test organisms;
(3) use of reference toxicants, controls, and
exposure replicates; (4) instrument
calibration; (5) record keeping; and (6) data
evaluation. For other types of ecological
assessments, however, QA/QC procedures
are less well defined (e.g., for biosurveys of
vegetation, terrestrial vertebrates). BTAG
HIGHLIGHT 4-3
Elements of a QAPP
(1) Project description
(2) Designation of QA/QC
responsibilities
(3) Statistical tests and data quality
objectives
(4) Sample collection and chain of
custody
(5) Sample analysis
(6) System controls and preventive
maintenance
(7) Record keeping
(8) Audits
(9) Corrective actions
(10) Quality control reports
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members can provide input on appropriate QA/QC procedures based on their experience with
Superfund sites.
4.4.3 Field Verification of Sampling Plan and Contingency Plans
For biological sampling, uncontrolled variables can influence the availability of species
to be sampled, the efficiency of different types of sampling techniques, and the level of effort
required to achieve the sample sizes specified in the SAP. As a consequence, the risk
assessor should develop a plan to test the sampling design before the WP and SAP are signed
and the •site investigation begins. Otherwise, field sampling during the site investigation could
fail to meet the DQOs specified in the SAP, and the study could fail to meet its objectives.
Step 5 provides a description of the field verification of the sampling design.
To the extent that potential field problems can be anticipated, contingency plans also
should be specified in the SAP. An example of a contingency plan is provided in Steps 5 and
6 (Examples 5-2 and 6-1).
4.5 SCIENTIFIC/MANAGEMENT DECISION POINT (SMDP)
The completion of the ecological risk assessment WP and SAP should coincide with
an SMDP. Within this SMDP, the ecological risk assessor and the ecological risk manager
agree on: (1) selection of measurement endpoints; (2) selection of the site investigation
methods; and (3) selection of data reduction and interpretation techniques. The WP or SAP
also should specify how inferences will be drawn from the measurement to the assessment
endpoints.
4.6 SUMMARY
At the conclusion of Step 4, there will be an agreement on the contents of the WP and
SAP. As noted earlier, these plans can be parts of a larger WP and SAP that are developed
to meet other remedial investigation needs, or they can be separate documents. When
possible, any field sampling efforts for the ecological risk assessment should overlap with
other site data collection efforts to reduce sampling costs and to prevent redundant sampling.
The WP and/or the SAP should specify the methods by which the collected data will
be analyzed. The plan(s) should include all food-chain-exposure-model parameters, data
reduction techniques, data interpretation methods, and statistical analyses that will be used.
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STEP 5: FIELD VERIFICATION OF SAMPLING DESIGN
OVERVIEW
4
Before the WP and SAP are signed, it is important to verify that the field
sampling plan they specify is appropriate and implementable at the site. If this has not
already been done, it should be done now. During field verification of the sampling
design, the testable hypotheses, exposure pathway models, and measurement endpoints
are evaluated for their appropriateness and implementability. The assessment
endpoint(s), however, should not be under evaluation in this step; the appropriateness
of the assessment endpoint should have been resolved in Step 3. If an assessment
endpoint is changed at this step, the risk assessor must return to Step 3, because the
enure process leading to the actual site investigation in Step 6 assumes the selection of
appropriate assessment endpoints.
5.1 PURPOSE
The primary purpose of field verification of the sampling plan is to ensure that the
samples specified by the SAP actually can be collected. A species that will be associated
with a measurement endpoint and/or exposure point concentration should have been observed
at the preliminary site characterization or noted during previous site visits. During this step,
previously obtained information should be verified and the feasibility of sampling will need to
be checked by a site visit. Preliminary sampling will determine if the targeted species is
present and—equally important—collectable in sufficient numbers or total biomass to meet
data quality objectives. This preliminary field assessment also allows for final confirmation
of the habitats that exist on or near the site. Habitat maps are verified a final time, and
interpretations of aerial photographs can be checked.
Final decisions on reference areas also should be made in this step. The reference
areas should be chosen to be as similar as possible to the site in all aspects except
contamination. Parameters to be evaluated for similarity include, but are not limited to:
slope, habitat, species potentially present, soil and sediment characteristics, and for surface
waters, flow rates, substrate type, water depth, temperature, turbidity, oxygen levels, water
hardness, pH, and other standard water quality parameters. If several on-site habitats or
habitat variables are being investigated, then several reference areas could be required.
Reference areas should be as free of site-related contaminants above background levels as
practical.
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5.2 DETERMINING SAMPLING FEASIBILITY
When sampling biota, it is difficult to predict what level of effort will be necessary to
obtain an adequate number of individuals of the required size. Some preliminary field
measurements often can help determine adequate sampling efforts to attain the sample sizes
specified in the SAP for statistical analyses. The WP and SAP should be signed and the site
investigation should be implemented immediately after verification of the sampling design to
limit effects of uncontrolled field variables. For example, evaluation of current small
mammal population density might indicate to the investigator that 400 trap-nights instead of
50 are necessary to collect the required number of small mammals. If there is a time lag
between the field sampling verification and the actual site investigation, it could be necessary
to reverify the field sampling to determine if conditions have changed.
Sampling methods for abiotic media also should be tested. There is a wide variety of
sampling devices and methods, and it is important to use the most appropriate, as the
following examples illustrate:
• When sampling a stream's surface water, if the stream is only three inches
deep, collecting the water directly into 32-ounce bottles would not be practical.
• Sampling the substrate in a stream might be desirable, but if the substrate is
bedrock, it might not be feasible or the intent of the sampling design.
An exposure-response relationship between contamination and biological effects is a
key component of establishing causality during the analysis phase of the baseline risk
assessment (Step 6). If extent-of-contamination sampling is conducted in phases, abiotic
exposure media and biotic samples must be collected simultaneously because the interactions
(both temporal and spatial) between the matrix to be remediated and the biota are crucial to
the development of a field exposure-response relationship. Failure to collect one sample
properly or to coordinate samples temporally can significantly impact the interpretation of the
data.
Sampling locations need to be checked to make sure that they are appropriately
described and placed within the context of the sampling plan. Directions for a sediment
sample "to be taken 5 feet from the north side of stream A," could cause confusion if the
stream is only 4 feet wide, or if the sampler doesn't know if the sample should be taken in
the stream, or 5 feet away from the edge of the stream. All samples should be checked
against the intended use of the data to be obtained.
All pathways for the migration of contaminants off site should be evaluated, such as
windblown dust, surface water runoff, and erosion. Along these pathways, a gradient of
decreasing contamination with increasing distance from the site might exist. Site-specific
ecological evaluations and risk assessments can be more useful to risk managers if gradients
of contamination can be located and evaluated.
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Contaminant migration pathways might have changed, either due to natural causes
(e.g., storms) or site remediation activities (e.g., erosion channels might have been filled or
dug up to prevent further migration of contaminants). Channels of small or large streams,
brooks, or rivers might have moved; sites might have been flooded. All of the assumptions
of the migration and exposure pathways need to be verified prior to the full site investigation.
If a contaminant gradient is necessary for the sampling plan, it is important to verify that the
gradient exists and that the range of contaminant concentrations is appropriate. A gradient of
contamination that causes no impacts at the highest concentration measured has as little value
as a gradient that kills everything at the lowest concentration measured; in either case, the
gradient would not provide useful exposure-response information. A gradient verification
requires chemical sampling, but field screening-level analyses might be effective.
These and other problems associated with the practical implementation of sampling
should be resolved prior to finalizing the SAP to the extent practicable. Assessing the
feasibility of the sampling plan before the site investigation begins saves costs in the long
term because it minimizes the chances of failing to meet DQOs during the site investigation.
Examples 5-1 and 5-2 describe the field verification of the sampling plan for the
hypothetical copper and DDT sites illustrated in Appendix A. Note that the scope of the field
verification differs for the copper and DDT sites. For the DDT site, a modification to the
study design was necessary. For both sites, the issues were resolved and a sign-off was
obtained at the SMDP for this step.
Any change in measurement endpoints will require that exposure pathways to the new
measurement endpoint be checked. The new measurement endpoint must fit into the
established conceptual model. Changes to measurement endpoints might require revision of
the conceptual model and agreement to the changes at the SMDP. It is highly desirable that
the agreed-upon conceptual model should be modified and approved by the same basic group
of individuals who developed it. .
5.3 SCIENTIFIC/MANAGEMENT DECISION POINT (SMDP)
The SMDP for the field verification of the sampling design is the signing of the
finalized WP and SAP. Any changes to the investigation proposed in Step 4 must be made
with agreement from the risk manager and risk assessment team. The risk manager must
understand what changes have been made and why, and must ensure that the risk management
decisions can be made from the information that the new study design can provide. The risk
assessors must be involved to ensure that the assessment endpoints and testable hypotheses
are still being addressed.
In the worst cases, changes in the measurement endpoints could be necessary, with
corresponding changes to the risk hypotheses and sampling design. Any new measurement
endpoints must be evaluated according to their utility for inferring changes in the assessment
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EXAMPLE 5-1
Field Verification of Sampling Design-Copper Site
Copper was released from a seep area of a landfill adjacent to a small pond; the release
and resulting elevated copper levels in the pond are of concern. The problem formulation and
conceptual model stated that the assessment endpoint was the maintenance of a typical pond
community for the area, including the benthic invertebrates and fish. Toxicity testing was
selected to evaluate the potential toxicity of copper to aquatic organisms. Three toxicity tests
were selected: a 10-day solid-phase sediment toxicity test (with the amphipod Hyalella azteca),
and two water column tests (i.e., the 7-day growth test with the green alga Selenastrum
capricornutum and the fathead minnow, Pimephales promelas, 1-day larval growth test). The
study design specified that sediment and water for the toxicity tests would be collected at the
leachate seeps known to be at the pond edge, and at three additional equidistant locations
transecting the pond (including the point of maximum pond depth). The pond contains water
year-round; however, the seep flow depends on rainfall. Therefore, it is only necessary to verify
that the leachate seep is active at the time of sampling.
endpoints and their compatibility with the site conceptual model (from Steps 3 and 4). Loss
of the relationship between measurement endpoints and the assessment endpoints, the risk
questions or testable hypothesis, and the site conceptual model will result in a failure to meet
study objectives.
Despite one's best efforts to conduct a sound site assessment, unexpected
circumstances might still make it necessary for the sampling plan to be changed in the field.
Any changes should be agreed to and documented by the lead risk assessor in consultation
with the risk manager.
Once the finalized WP and SAP are approved and signed, Step 6 should begin.
5.4 SUMMARY
In summary, field verification of the sampling plan is very important to ensuring that
the DQOs of the site investigation can be met. This step verifies that the selected assessment
endpoints, testable hypotheses, exposure pathway model, measurement endpoints, and study
design from Steps 3 and 4 are appropriate and implementable at the site. By verifying the
field sampling plan prior to conducting the full site investigation, well-considered alterations
can be made to the study design and/or implementation if necessary. These changes will
ensure that the ecological risk assessment meets the study objectives.
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If changing conditions force changes to the sampling plan in the field (e.g., selection
of a different reference site), the changes should be agreed to and documented by the lead
risk assessor in consultation with the risk manager.
EXAMPLE 5-2
Field Verification of Sampling Design-DDT Site
For the stream DDT site, the assessment endpoint was protection of piscivorous birds
from adverse reproductive effects. The conceptual model included the exposure pathway of
sediment to forage fish to the kingfisher. The measurement endpoint selected was tissue residue
' levels in creek chub (Semotilus atromaculatus), which could be associated with contaminant
levels in sediments. Existing information on the stream contamination indicates that a gradient
of contamination exists and that five specific sampling locations should be sufficient to
characterize the gradient to the point where concentrations are unlikely to have adverse effects.
The study design specified that 10 creek chub of the same size and sex be collected at each
location. Each chub should be approximately 20 grams, so that minimum sample mass
requirements could be met without using composite samples for analysis. In addition, QA/QC
protocol requires that 10 more fish be collected at one of the locations.
In this example, a site assessment was necessary to verify that a sufficient number of
creek chub of the specified size would be present to meet the sampling requirements. Stream
conditions were evaluated to determine what fish sampling technique would work at the targeted
locations. A field assessment was conducted, and several fish collection techniques were used
to determine which was the most effective for the site. Collected creek chub and other fish
were examined to determine the size range available and whether the sex of the individuals
could be determined. . . . .
The site assessment indicated that the creek chub might not be present in sufficient
numbers to provide the necessary biomass for chemical analyses. Based upon those findings, a
contingency plan was agreed to, which stated that both the creek chub and the longnosed dace
(Rhinichthys cataractae) would be collected. If the creek chub were collected at all locations in
sufficient numbers, then those samples would be analyzed and the dace would be released. If
sufficient creek chub could not be collected but sufficient longnosed dace could, the longnosed
dace would be analyzed and the creek chub released. If neither species could be collected at all
locations in sufficient numbers, then a mix of the two species would be used; however, for any
given sampling location only one species would be used to make the sample. In addition, at
one location, which preferably had high DDT levels in the sediment, sufficient numbers (20
grams) of both species would be collected to allow comparison (and calibration) of the
accumulation between the two species.
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STEP 6: SITE INVESTIGATION AND ANALYSIS PHASE
OVERVIEW
Information collected during the site investigation is used to characterize
exposures and ecological effects. The site investigation includes all of the field
sampling and surveys that are conducted as part of the ecological risk assessment. The
site investigation and analysis of exposure and effects should be straightforward,
following the WP and SAP developed in Step 4 and tested in Step 5.
Exposure characterization relies heavily on data from the site investigation and
can involve fate-and-transport modeling. Much of the information for characterizing
potential ecological effects was gathered from the literature review during problem
formulation, but the site investigation might provide evidence of existing ecological
impacts and additional exposure-response information.
6.1 INTRODUCTION
The site investigation (Section 6.2) and analysis phase (Section 6.3) of the ecological
risk assessment should be straightforward. In Step 4, all issues related to the study design,
sample collection, DQOs, and procedures for data reduction and interpretation should have
been identified and resolved. However, as described in Step 5, there are circumstances that
can arise during a site investigation that could require modifications to the original study
design. If any unforeseen events do require a change to the WP or SAP, all changes must be
agreed upon at the SMDP (Section 6.4). The results of Step 6 are used to characterize
ecological risks in Step 7.
6.2 SITE INVESTIGATION
The WP for the site investigation is based on the site conceptual model and should
specify the assessment endpoints, risk questions, and testable hypotheses. The SAP for the
site investigation should specify the relationship between measurement and assessment
endpoints, the necessary number, volume, and types of samples to be collected, and the
sampling techniques to be used. The SAP also should specify the data reduction and
interpretation techniques and the DQOs. The feasibility of the sampling design was tested in
Step 5. Therefore, the site investigation should be a direct implementation of the previously
designed study.
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During the site investigation, it is important to adhere to the DQOs and to any
requirements for co-located sampling. Failure to collect one sample properly or to coordinate
samples temporally can significantly affect interpretation of ihe data. Changing field
conditions (Section 6.2.1) and new information on the nature and extent of contamination
(Section 6.2.2) can require a change in the SAP.
6.2.1 Changing Field Conditions
In instances where unexpected conditions arise in the field that make the collection of
specified samples impractical or not ideal, the ecological risk assessor should reevaluate the
feasibility of the sampling design as described in Step 5. Field efforts should not necessarily
be halted, but decisions to change sampling procedures or design must be agreed to by the
risk manager and lead risk assessor or project-delegated equivalents.
Field modifications to study designs are not uncommon during field investigations.
When the WP and SAP provide a precise conceptual model and study design- with specified
data analyses, informed modifications to the SAP can be made to comply with the objectives
of the study. As indicated in Step 4, contingency plans can be included in the original SAP
in anticipation of situations that might arise during the site investigation (see Example 6-1).
Any modifications, and the reasons for the modifications, must be documented in the baseline
risk assessment.
EXAMPLE 6-1
Fish Sampling Contingency Plan-DDT Site
At the DDT site where creek chub are to be collected for DDT tissue residue analyses,
a contingency plan for the site investigation was developed. An alternate species, the longnosed
dace, was specified with the expectation that, at one or all locations, the creek chub might be
absent at the time of the site investigation. Such contingency plans are prudent even when the
verification of the field sampling design described in Step 5 indicates that the samples are
obtainable.
6.2.2 Unexpected Nature or Extent of Contamination
*.
It is not uncommon for an initial sampling phase of the RI to reveal that
contamination at levels of concern extend beyond areas initially established for characterizing
contamination and ecological effects at the site or that contaminant gradients are much steeper
than anticipated. If this contingency changes the opportunity for evaluating biological effects
along a contamination gradient, the ecological risk assessors and risk manager need to
determine whether additional sampling (e.g., further downstream from the site) is needed.
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Thus, it is important for the ecological risk assessors to track information on the nature and
extent of contamination as RI sampling is conducted.
On occasion, new contaminants are identified during an RI. In this case, the risk
assessors and site manager will need to return to Step 1 to screen the new contaminants for
ecological risk.
Immediate analysis of the data for each type of sampling and communication between
the risk assessors and risk managers can help ensure that the site investigation is adequate to
achieve the study goals and objectives when field modifications are necessary. If a change to
the WP or SAP is needed, the lead risk assessor and risk manager must agree on all changes
(the SMDP in Section 6.4).
6.3 ANALYSIS OF ECOLOGICAL EXPOSURES AND EFFECTS
The analysis phase of the ecological risk assessment consists of the technical
evaluation of data on existing and potential exposures (Section 6.3.1) and ecological effects
(Section 6.3.2) at the site. The analysis is based on the information collected during Steps 1
'through 5 and often includes additional assumptions or models to interpret the data in the
context of the site conceptual model. As illustrated in Exhibit 6-1, analysis of exposure and
effects is performed interactively, with the analysis of one informing the analysis of the other.
This step follows the data interpretation and analysis methods specified in the WP and SAP,
and therefore should be a straightforward process.
In the analysis phase, the site-specific data obtained during the site investigation
replace many of the assumptions that were made for the screening-level analysis in Steps 1
and 2. For the exposure and ecological effects characterizations, the uncertainties associated
with the field measurements and with assumptions where site-specific data are not available
must be documented.
6.3.1 Characterizing Exposures
Exposure can be expressed as the co-occurrence or contact of the stressor with the
ecological components, both in time and space (U.S. EPA, 1992a). Thus, both the stressor
and the ecosystem must be characterized on similar temporal and spatial scales. The result of
the exposure analysis is an exposure profile that quantifies the magnitude and spatial and
temporal patterns of exposure as they relate to the assessment endpoints and risk questions
developed during problem formulation. The exposure profile and a description of associated
uncertainties and assumptions serve as input to the risk characterization in Step 7.
Stressor characterization involves determining the stressor's distribution and pattern of
change. The analytic approach for characterizing ecological exposures should have been
established in the WP and SAP on the basis of the site conceptual model. For chemical
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EXHIBIT 6-1
Analysis Phase (U.S. EPA, 1992a)
PROBLEM FORMULATION
ANALYSIS
v RISK CHARACTERIZATION
N
PROBLEM FORMULATION
Characterization of Exposure
•
Stressor x Ecosy
Pattern of Change Abie
\\ //
Y Exposure [/
N Analysis
1
Exposure
Profile
IHLV
Characterization of
stem
tic
\\
Ecological Effects
^
Evaluation
of Relevant
Effects Data
./
\ Ecological
Response
Analysis
1
Stressor-Response
Profile
TaT
•M^^B^^M
7
••••••••
o
o>
ST
B>
£
sr
§
o*
5"
a
§
o
9
ID
RISK CHARACTERIZATION
6-4
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stressors at Superfund sites, usually a combination of fate-and-transport modeling and
sampling data from the site are used to predict the current and likely future nature and extent
of contamination at a site.
HIGHLIGHT 6-1
Uncertainty in Exposure Models
The accuracy of an exposure model
depends on the accuracy of the input
parameter values and the validity of the
model's structure (i.e., the degree to which it
represents the actual relationships among
parameters at the site). Field measurements
can be used to calibrate model outputs or
intermediate calculations. Such field
measurements should be specified in the WP
and SAP. For example, studies of tissue
residue levels often are used to calibrate
exposure and food-chain models.
When characterizing exposures, the
ecological context of the site established
during problem formulation is analyzed
further, both to understand potential effects
of the ecosystem on fate and transport of
chemicals in the environment and to
evaluate site-specific characteristics of
species or communities of concern. Any
site-specific information that can be used to
replace assumptions based on information
from the literature or from other sites is
incorporated into the description of the
ecological components of the site.
Remaining assumptions and uncertainties in
the exposure model (Highlight 6-1) should
be documented.
6.3.2 Characterizing Ecological Effects
At this point, all evidence for existing and potential adverse effects on the assessment
endpoints is analyzed. The information from the literature review on ecological effects is
integrated with any evidence of existing impacts based on the site investigation (e.g., toxicity
testing). The methods for analyzing site-specific data should have been specified in the WP
and SAP, and thus should be straightforward. Both exposure-response information and
evidence that site contaminants are causing or can cause adverse effects are evaluated.
Exposure-response analysis. The exposure-response analysis for a Superfund site
describes the relationship between the magnitude, frequency, or duration of a contaminant
stressor in an experimental or observational setting and the magnitude of response. In this
phase of the analysis, measurement endpoints are related to the assessment endpoints using
the logical structure provided by the conceptual model. Any extrapolations that are required
to relate measurement to assessment endpoints (e.g., between species, between response
levels, from laboratory to field) are explained. Finally, an exposure-response relationship is
described to the extent possible (e.g., by a regression equation), including the confidence
limits (quantitative or qualitative) associated with the relationship.
Under some circumstances, site-specific exposure-response information can be
obtained by evaluating existing ecological impacts along a contamination gradient at the site.
Statistical techniques to identify or describe the relationship between exposure and response
from the field data should have been specified in the WP and SAP. The potential for
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confounding stressors that might correlate with the contamination gradient should be
documented (e.g., decreasing water temperature downstream of a site; reduced soil erosion
further from a site).
An exposure-response analysis is of particular importance to risk managers who must
balance human health and ecological concerns against the feasibility and effectiveness of
remedial options. An exposure-response function can help a risk manager to specify the
trade-off between the degree of cleanup and likely benefits of the cleanup and to balance
ecological and financial costs and benefits of different remedial options, as discussed in
Step 8.
When exposure-response data are not available or cannot be developed, a threshold for
adverse effects can be developed instead, as in Step 2. For the baseline risk assessment,
however, site-specific information should be used instead of conservative assumptions
whenever possible.
Evidence of causality. At Superfund sites, evidence of causality is key to the risk
assessment. Thus, it is important to evaluate the strength of the causal association between
site-related contaminants and effects on the measurement and assessment endpoints.
Demonstrating a correlation between a contaminant gradient and ecological impacts at a site
is a key component of establishing causality, but other evidence can be used in the absence of
such a demonstration. Moreover, an exposure-response correlation at a site is not sufficient to
demonstrate causality, but requires one or more types of supporting evidence and analysis of
potential confounding factors. Hill's (1965) criteria for evaluating causal associations are
outlined in the Framework (U.S. EPA, 1992a).
6.4 SCIENTIFIC/MANAGEMENT DECISION POINT (SMDP)
An SMDP during the site investigation and analysis phase is needed only if alterations
to the WP or SAP become necessary. In the worst case, changes in measurement endpoints
could be required, with corresponding changes to the testable hypotheses and sampling
design. Any new measurement endpoints must be evaluated according to their utility for
inferring changes in the assessment endpoints and their compatibility with the site conceptual
model; otherwise, the study could fail to meet its objectives.
Proposed changes to the SAP must be made in consultation with the risk manager and
the risk assessors. The risk manager must understand what changes have been made and
why, and must ensure that the risk management decisions can be made from the information
that the new study design can provide. The risk assessors must be involved to ensure that the
assessment endpoints and study questions or testable hypotheses are still being addressed.
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6.5 SUMMARY
The site investigation step of the ecological risk assessment should be a
straightforward implementation of the study designed in Step 4 and verified in Step 5. In
instances where unexpected conditions arise in the field that indicate a need to change the
study design, the ecological risk assessors should reevaluate the feasibility or adequacy of the
sampling design. Any proposed changes to the WP or SAP must be agreed upon by both the
risk assessment team and the risk manager and must be documented in the baseline risk
assessment.
The analysis phase of the ecological risk assessment consists of the technical
evaluation of data on existing and potential exposures and ecological effects and is based on
the information collected during Steps 1 through 5 and the site investigation in Step 6.
Analyses of exposure and effects are performed interactively, and follow the data
interpretation and analysis methods specified in the WP and SAP. Site-specific data obtained
during Step 6 replace many of the assumptions that were made for the screening-level
analysis in Steps 1 and 2. Evidence of an exposure-response relationship between
contamination and ecological responses at a site helps to establish causality. The results of
Step 6 are used to characterize ecological risks in Step 7.
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STEP 7: RISK CHARACTERIZATION
OVERVIEW
In risk characterization, data on exposure and effects are integrated into a
statement about risk to the assessment endpoints established during problem
formulation. A weight-of-evidence approach is used to interpret the implications of
different studies or tests for the assessment endpoints. In a well-designed study, risk'
characterization should be straightforward, because the procedures were established in
the WP and SAP. The risk characterization section of the baseline ecological risk
assessment should include a qualitative and quantitative presentation of the risk results
and associated uncertainties.
7.1 INTRODUCTION
Risk characterization is the final phase of the risk assessment process and includes two
major components: risk estimation and risk description (U.S. EPA, 1992a; Exhibit 7-1). Risk
estimation (Section 7.2) consists of integrating the exposure profiles with the exposure-effects
information and summarizing the associated uncertainties. The risk description (Section 7.3)
provides information important for interpreting the risk results and, in the Superfund Program,
identifies a threshold for adverse effects on the assessment endpoints (Section 7.4).
It is U.S. EPA policy that risk characterization should be consistent with the values of
"transparency, clarity, consistency, and reasonableness" (U.S. EPA, 1995f). "Well-balanced
risk characterizations present risk conclusions and information regarding the strengths and
limitations of the assessment for other risk assessors, EPA decision-makers, and the public"
(U.S. EPA, 1995f). Thus, when preparing the risk characterization, the risk assessment team
should make sure that the documentation of risks is easy to follow and understand, with all
assumptions, defaults, uncertainties, professional judgments, and any other inputs to the risk
estimate clearly identified and easy to find.
7.2 RISK ESTIMATION
Documentation of the risk estimates should describe how inferences are made from the
measurement endpoints to the assessment endpoints established in problem formulation. As
stated earlier, it is not the purpose of this document to provide a detailed guidance on the
selection and utilization of risk models. The risk assessment team should have developed and
the risk manager should have agreed upon the conceptual model used to characterize risk, its
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EXHIBIT 7-1
Risk Characterization (U.S. EPA, 1992a)
PROBLEM FORMULATION
ANALYSIS
x RISK CHARACTERIZATION
X
ANALYSIS
1
Risk Estimation
Integration
Uncertainty
Analysis
J
Ecological
Risk
Summary
Interpretation of
Ecological
Significance
I
Discussion Between the
Risk Assessor and Risk Manager
(Results)
s?
2.
5*
o
0
(O
Risk Management
7-2
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assumptions, uncertainties, and interpretation in Steps 3 through 5. This agreement is
specified in the site WP and SAP and is the purpose of the SMDPs in Steps 3 through 5.
Unless the site investigation during Step 6 discovers unexpected information, the risk
assessment should move smoothly through the risk characterization phase, because the data
interpretation procedures were specified in the WP and SAP. While it might be informative
to investigate a data set for trends, outliers, or other statistical indicators, these investigations
should be secondary to the data interpretations specified in the SAP. Analysis of the data
beyond the purposes for which it was collected might be informative, but could lead to
biased, conflicting, or superfluous conclusions. Those outcomes can divert or confound the
risk characterization process.
For ecological risk assessments that entail more than one type of study (or line of
evidence), a strength-of-evidence approach is used to integrate different types of data to
support a conclusion. The data might include toxicity test results, assessments of existing
impacts at a site, or risk calculations comparing exposures estimated for the site with toxicity
values from the literature. Balancing and interpreting the different types of data can be a
major task and require professional judgment. As indicated above, the strength of evidence
provided by different types of tests and the precedence that one type of study might have over
another should already have been established during Step 4. Taking this approach will ensure
that data interpretation is objective and not biased to support a preconceived answer.
Additional strength-of-evidence considerations at this stage include the degree to which DQOs
were met and whether confounding factors became evident during the site investigation and
analysis phase.
For some biological tests (e.g., toxicity tests, benthic macroinvertebrate studies), all or
some of the data interpretation process is outlined in existing documents, such as in toxicity
testing manuals. However, in most cases, the SAP must provide the details on how the data
are to be interpreted for a site. The data interpretation methods also should be presented in
the risk characterization documentation. For example, if the triad approach was used to
evaluate contaminated sediments, the risk estimation section should describe how the three
types of studies (i.e., toxicity test, benthic invertebrate survey, and sediment chemistry) are
integrated to draw conclusions about risk.
Where exposure-response functions are not available or developed, the quotient
method of comparing an estimated exposure concentration to a threshold for response can be
used, as in Step 2. Whenever possible, however, presentation of full exposure-response
functions provides the risk manager with more information on which to base site decisions.
This guidance has recommended the use of on-site contamination gradients to demonstrate on-
site exposure-response functions. Where such data have been collected, they should be
presented along with the risk estimates. Hazard quotients, hazard indices (for contaminants
with the same mechanism of toxicity), the results of in situ toxicity testing, or community
survey data can be mapped along with analytic chemistry data to provide a clear picture of
the relationship between areas of contamination and effects.
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In addition to developing point estimates of exposure concentrations, as for the hazard
quotient approach, it might be possible to develop a distribution of exposure levels based on
the potential variability in various exposure parameters (see Section 7.3.2). Probabilities of
exceeding a threshold for adverse effects might then be estimated. Again, the risk assessment
team and risk manager should have already agreed to what analyses will be used to
characterize risks.
7.3 RISK DESCRIPTION
A key to risk description for Superfund sites is documentation of environmental
contamination levels that bound the threshold for adverse effects on the assessment endpoints
(Section 7.3.1). The risk description can also provide information to help the risk manager
judge the likelihood and ecological significance of the estimated risks (Sections 7.3.2 and
7.3.3, respectively).
7.3.1 Threshold for Effects on Assessment Endpoints
Key outputs of the risk characterization step are contaminant concentrations in each
environmental medium that bound the threshold for estimated adverse ecological effects given
the uncertainty inherent in the data and models used. The lower bound of the threshold
would be based on consistent conservative assumptions and NOAEL toxicity values. The
upper bound would be based on observed impacts or predictions that ecological impacts could
occur. This upper bound would be developed using consistent assumptions, site-specific data,
LOAEL toxicity values, or an impact evaluation.
The approach to estimating environmental contaminant concentrations that represent
thresholds for adverse ecological effects should have been specified in the study design (Step
4); When higher-trophic-level organisms are associated with assessment endpoints, the study
design,should have described how monitoring data and contaminant-transfer models would be
used to back-calculate an environmental concentration representing a threshold for effect. If
the site investigation demonstrated a gradient of ecological effects along a contamination
gradient, the risk assessment team can identify and document the levels of contamination
below which no further improvements in the assessment endpoints are discernable or
expected. If departures from the original analysis plan are necessary based on information
obtained during the site investigation or data analysis phase, the reasons for change should be
documented.
When assessment endpoints include populations of animals that can travel moderate
distances, different ways of presenting a threshold for adverse effects are possible. Various
combinations of level of contamination and area! extent of contamination relative to the
foraging range of the animals can result in similar contaminant intake levels by the animals.
In that case, a point of departure for identifying a threshold for effect would be to identify
that level of contamination, which if uniformly distributed both at the site and beyond, would
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not pose a threat. The assumption of uniform contamination has been used to back-calculate
water-quality criteria to protect piscivorous wildlife in the Great Lakes (U.S. EPA, 1995a).
Again, use of this approach should have been specified in the study design.
7.3.2 Likelihood of Risk
In addition to identifying one or more thresholds for effects, the risk assessment team
might develop estimates of the probability that exposure levels would exceed the ecotoxicity
thresholds given the distribution of values likely for various exposure parameters (e.g., home
range size, population density). A distributional analysis might be used to estimate the range
of likely exposure levels associated with a given exposure model based on ranges for the
input variables.
7.3.3 Additional Risk Information
In addition to developing numerical estimates of existing impacts, risks, and thresholds
for effect, the risk assessor should put the estimates in context with a description of their
extent, magnitude, and potential ecological significance. Additional ecological risk
descriptors are listed below:
• The location and area] extent of existing contamination above a threshold for
adverse effects;
• The degree to which the threshold for contamination is exceeded or is likely to
be exceeded in the future, particularly if exposure-response functions are
available; and
• The expected half-life (qualitative or quantitative) of contaminants in the
environment (e.g., sediments, food chain) and the potential for natural recovery
once the sources of contamination are removed.
To interpret the information in light of remedial options, the risk manager might need to
solicit input from specific experts.
At this stage, it is important for the risk assessors to consider carefully several
principles of risk communication, as described in U.S. EPA's (1996a) Proposed Guidelines
for Ecological Risk Assessment.
7.4 UNCERTAINTY ANALYSIS
There are several sources of uncertainties associated with Superfund ecological risk
estimates. One is the initial selection of substances of concern based on the sampling data .
and available toxicity information. Other sources of uncertainty include estimates of toxicity
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to ecological receptors at the site based on limited data from the laboratory (usually on other
species), from other ecosystems, or from the site over a limited period of time. Additional
uncertainties result from the exposure assessment, as a consequence of the uncertainty in
chemical monitoring data and models used to estimate exposure concentrations or doses.
Finally, further uncertainties are included in risk estimates when simultaneous exposures to
multiple substances occur.
Uncertainty should be distinguished from variability, which arises from true
heterogeneity or variation in characteristics of the environment and receptors. Uncertainty, on
the other hand, represents lack of knowledge about certain factors which can sometimes be
reduced by additional study.
This section briefly notes several categories of uncertainty (Section 7.4.1) and
techniques for tracking uncertainty through a risk assessment (Section 7.4.2). Additional
guidance on discussing uncertainty and variability in risk characterization is provided in U.S.
EPA's (19920 Guidance on Risk Characterization for Risk Managers and Risk Assessors.
7A.I Categories of Uncertainty
There are three basic categories of uncertainties that apply to Superfund site risk
assessments: (1) conceptual model uncertainties; (2) natural variation and parameter error; and
(3) model error. Each of these is described below.
There will be uncertainties associated with the conceptual model used as the basis to
investigate the site. The initial characterization of the ecological problems at a Superfund
site, likely exposure pathways, chemicals of concern, and exposed ecological components,
requires professional judgments and assumptions. To the extent possible, the risk assessment
team should describe what judgments and assumptions were included in the conceptual model
that formed the basis of the WP and SAP.
Parameter values (e.g., water concentrations, tissue residue levels, food ingestion rates)
usually can be characterized as a distribution of values, described by central tendencies,
ranges, and percentiles, among other descriptors. When evaluating uncertainty in parameter
values, it is important to distinguish uncertainty from variability. Ecosystems include highly
variable abiotic (e.g., weather, soils) and biotic (e.g., population density) components. If all
instances of a parameter (e.g., all members of a population) could be sampled, the "true"
parameter value distribution could be described. In practical terms, however, only a fraction
of the instances (e.g., a few of the members of the population) can be sampled, leaving
uncertainty concerning the true parameter value distribution. The risk assessor should provide
either quantitative or qualitative descriptions of uncertainties in parameter value distributions.
Finally, there is uncertainty associated with how well a model (e.g., fate and transport
model) approximates true relationships between site-specific environmental conditions.
Models available at present tend to be fairly simple and at best, only partially validated with
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field tests. As a consequence, it is important to identify key model assumptions and their
potential impacts on the risk estimates.
7.4.2 Tracking Uncertainties
In general, there are two approaches to tracking uncertainties through a risk
assessment: (1) using various point estimates of exposure and response to develop one or
more point estimates of risk; and (2) conducting a distributional analysis to predict a
distribution of risks based on a distribution of exposure levels and exposure-response
information. Whether one or the other or both approaches are taken should have been agreed
to during Step 4, and the specific type of analyses to be conducted should have been specified
in the SAP.
7.5 SUMMARY
Risk characterization integrates the results of the exposure profile and exposure-
response analyses, and is the final phase of the risk assessment process. It consists of risk
estimation and risk description, which together provide information to help judge the
ecological significance of risk estimates in the absence of remedial activities. The risk
description also identifies a threshold for effects on the assessment endpoint as a range
between contamination levels identified as posing no ecological risk and the lowest
contamination levels identified as likely to produce adverse ecological effects. To ensure that
the risk characterization is transparent, clear, and reasonable, information regarding the
strengths and limitations of the assessment must be identified and described.
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STEPS: RISK MANAGEMENT
OVERVIEW
Risk management at a Superfund site is ultimately the responsibility of the site
risk manager, who must balance risk reductions associated with cleanup of
contaminants with potential impacts of the remedial actions themselves. The risk
manager considers inputs from the risk assessors, BTAGs, stakeholders, and other
involved parties. In Step 7, the risk assessment team identified a threshold for effects
on the assessment endpoint as a range between contamination levels identified as
posing no ecological risk and the lowest contamination levels identified as likely to
produce adverse ecological effects. In Step 8, the risk manager evaluates several
factors in deciding whether or not to clean up to within that range.
8.1 INTRODUCTION
Risk management is a distinctly different process from risk assessment (NRC, 1983,
1994; U.S. EPA, 1984a, 1995f). The risk assessment establishes whether a risk is present and
defines a range or magnitude of the risk. In risk management, the results of the risk
assessment are integrated with other considerations to make and justify risk management
decisions. Additional risk management considerations can include the implications of existing
background levels of contamination, available technologies, tradeoffs between human and
ecological concerns, costs of alternative actions, and remedy selection. For further
information on management of ecological risks Agency-wide, see U.S. EPA 1994h. Some
Superfund-specific considerations are described below.
8.2 ECOLOGICAL RISK MANAGEMENT IN SUPERFUND
»
According to section 300.40 of the NCP, the purpose of the remedy selection process
is to eliminate, reduce, or control risks to human health and the environment. The NCP
indicates further that the results of the baseline risk assessment will help to establish
acceptable exposure levels for use in developing remedial alternatives during the FS. Based
on the criteria for selecting the preferred remedy and, using information from the human
health and ecological risk assessments and the evaluation of remedial options in the FS, the
risk .manager then selects a preferred remedy.
The risk manager must consider several types of information in addition to the
baseline ecological risk assessment when evaluating remedial options (Section 8.2.1). Of
8-1
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particular concern for ecological risk management at Superfund sites is the potential for
remedial actions themselves to cause adverse ecological impacts (Section 8.2.2). There also
exists the opportunity to monitor ecological components at the site to gauge the effectiveness
(or impacts) of the selected remedy (Section 8.2.3).
8.2.1 Other Risk Management Considerations
The baseline ecological risk assessment is not the only set of information that the risk
manager must consider when evaluating remedial options during the FS phase of the
Superfund process. The NCP (40 CFR 300.430(f)(l)(i)) specifies that each remedial
alternative should be evaluated according to nine criteria. Two are considered threshold
criteria, and take precedence over the others:
(1) Overall protection of human health and the environment; and
(2) Compliance with applicable or relevant and appropriate requirements (ARARs)
(unless waiver applicable).
As described in Section 8.2.2 below, a particularly important consideration for the first
criterion are the ecological impacts of the remedial options.
Five of the nine criteria are considered primary balancing criteria to be considered
after the threshold criteria:
(3) Long-term effectiveness and permanence;
(4) Reduction of toxicity, mobility, or volume of hazardous wastes through the use
of treatment;
(5) Short-term effectiveness;
(6) Implementability; and
(7) Cost.
Finally, two additional criteria are referred to as modifying criteria that must be
considered:
(8) State acceptance, and
. (9) Community acceptance.
Effective risk communication is particularly important to help ensure that a remedial option
that best satisfies the other criteria can be implemented at a site. U.S. EPA's (1996a)
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Proposed Guidelines for Ecological Risk Assessment provides an overview of this topic and
identifies some of the relevant literature.
Additional factors that the site risk manager takes into consideration include existing
background levels (see U.S. EPA, 1994g); current and likely future land uses (see U.S. EPA,
1995c); current and likely future resource uses in the area; and local, regional, and national
ecological significance of the site. Consideration of the ecological impacts of remedial
options and residual risks associated with leaving contaminants in place are very important
considerations, as described in the next section.
8.2.2 Ecological Impacts of Remedial Options
Management of ecological risks must take into account the potential for impacts to the
ecological assessment endpoints from implementation of various remedial options. The risk
manager must balance: (1) residual risks posed by site contaminants before and after
implementation of the selected remedy with (2) the potential impacts of the selected remedy
on the environment independent of contaminant effects. The selection of a remedial
alternative could require tradeoffs between long-term and short-term risk.
The ecological risks posed by the "no action" alternative are the risks estimated by the
baseline ecological risk assessment. In addition, each remedial option is likely to have its
own ecological impact. This impact could be anything from a short-term loss to complete
and permanent loss of the present habitat and ecological communities. In instances where
substantial ecological impacts will result from the remedy (e.g., dredging a wetland), the risk
manager will need to consider ways to mitigate the impacts of the remedy and compare the
mitigated impacts to the threats posed by the site contamination.
During the FS, the boundaries of potential risk under the no-action alternative (i.e.,
baseline conditions) can be compared with the evaluation of potential impacts of the remedial
options to help justify the preferred remedy. As indicated above, the preferred remedy should
minimize the risk of long-term impacts that could result from the remedy and any residual
contamination. When the selected remedial option leaves some site contaminants presumed to
pose an ecological risk in place, the justification for the selected remedy must be clearly
documented.
In short, consideration of the environmental effects of the remedy itself might result in
a decision to allow contaminants to remain on site at levels higher than the threshold for
effects on the assessment endpoint. Thus, selection of the most appropriate ecologically
based remedy can result in residual contamination that presents some risk.
8.2.3 Monitoring
Ecological risk assessment is a relatively new field with limited data available to
validate its predictions. At sites where remedial actions are taken to reduce ecological
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impacts and risks, the results of the remediation efforts should be compared with the
predictions made during the ecological risk assessment.
While it often is difficult to demonstrate the effectiveness of remedial actions in
reducing human health risks, it often is possible to demonstrate the effectiveness of
remediations to reduce ecological risks, particularly if a several-year monitoring program is
established. The site conceptual model provides the conceptual basis for monitoring options,
and the site investigation should have indicated which options might be most practical for the
site. Monitoring also is important to assess the effectiveness of a no-action alternative. For
example, monitoring sediment contamination and benthic communities at intervals following
removal of a contaminant source allows one to test predictions of the potential for the
ecosystem to recover naturally over time.
8.3 SCIENTIFIC/MANAGEMENT DECISION POINT (SMDP)
The risk management decision is finalized in the Record of Decision (ROD). The
decision should minimize the risk of long-term impacts that could result from the remedy and
any residual contamination. When the selected remedy leaves residual contamination at levels
higher than the upper-bound estimate of the threshold for adverse effects on the assessment
endpoint, the risk manager should justify the decision (e.g., describe how a more complete
physical remedy could jeopardize an ecological community more than the residual
contamination).
8.4 SUMMARY
Risk-management decisions are the responsibility of the risk manager (the site
manager), not the risk assessor. The risk manager should have been involved in planning the
risk assessment; knowing the options available for reducing risks, the risk manager can help
to frame questions during the problem-formulation phase of the risk assessment.
The risk manager must understand the risk assessment, including its uncertainties,
assumptions, and level of resolution. With an understanding of potential adverse effects
posed by residual levels of site contaminants and posed by the remedial actions themselves,
the risk manager can balance the ecological costs and benefits of the available remedial
options. Understanding the uncertainties associated with the risk assessment also is critical to
evaluating the overall protectiveness of any remedy.
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BIBLIOGRAPHY
This combined reference list and bibliography is intended to provide a broad, but not
all inclusive, list of other materials that may provide useful information for ecological risk
assessments at Superfund sites. These documents include other Superfund Program guidance
documents, standard guides for toxicity testing, other EPA program office references with
potential applications at Superfund sites, and other ecological risk assessment reference
materials. References cited in the text are marked with an asterisk (*).
American Public Health Association (APHA). 1989. Standard Methods for Examination of
Water and Wastewater. 17th edition. Washington, DC: APHA.
American Society for Testing and Materials (ASTM). I994a. Annual Book of ASTM
Standards. Philadelphia, PA: ASTM.
/
American Society for Testing and Materials (ASTM). 1994b. Standard guide for conducting
sediment toxicity tests with freshwater invertebrates: ASTM Standard E 1383-94.
American Society for Testing and Materials (ASTM). 1993a. Standard terminology relating
to biological effects and environmental fate: ASTM Standard E 943-93.
American Society for Testing and Materials (ASTM). I993b. Standard guide for designing
biological tests with sediments: ASTM Standard El 525-93.
American Society for Testing and Materials (ASTM). 1993. ASTM Standards of Aquatic
Toxicology and Hazard Evaluation. . Philadelphia, PA: ASTM.
American Society for Testing and Materials (ASTM). 1992. Standard guide for conducting
sediment toxicity tests with freshwater invertebrates: ASTM Standard E 1383-92.
American Society for Testing and Materials (ASTM). 1992. Standard guide for conducting
10-day static sediment toxicity tests with marine and estuarine amphipods: ASTM
Standard E 1367-92.
American Society for Testing and Materials (ASTM). 1990. Standard guide for collection,
storage, characterization, and manipulation of sediments for lexicological testing:
ASTM Standard E 1391-90.
American Society for Testing and Materials (ASTM). 1988. Standard guide for conducting
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Alternatives at Hazardous Waste Sites and Spills. Washington, DC: Office of Solid
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GLOSSARY
This glossary includes definitions from several sources. A superscript number next to a
word identifies the reference from which the definition was adapted (listed at the end of the
Glossary).
Abiotic.1 Characterized by absence of life; abiotic materials include non-living environmental
media (e.g., water, soils, sediments); abiotic characteristics include such factors as light,
temperature, pH, humidity, and other physical and chemical influences.
Absorption Efficiency. A measure of the proportion of a substance that a living organism
absorbs across exchange boundaries (e.g., gastrointestinal tract).
Absorbed Dose.2 The amount of a substance penetrating the exchange boundaries of an
organism after contact. Absorbed dose for the inhalation and ingestion routes of exposure is
calculated from the intake and the absorption efficiency. Absorbed dose for dermal contact
depends on the surface area exposed and absorption efficiency.
Accuracy.4 The degree to which a measurement reflects the true value of a variable.
Acute. Having a sudden onset or lasting a short time. An acute stimulus is severe enough
to induce a response rapidly. The word acute can be used to define either the exposure or the
response to an exposure (effect). The duration of an acute aquatic toxicity test is generally 4
days of less and mortality is the response usually measured.
Acute Response. The response of (effect on) an organisms which has a rapid onset. A
commonly measured rapid-onset response in toxicity tests is mortality.
i
Acute Tests. A toxicity test of short duration, typically 4 days or less (i.e., of short duration
relative to the lifespan of the test organism).
Administered Dose.2 The mass of a substance given to an organism and in contact with an
exchange boundary (i.e., gastrointestinal tract) per unit wet body weight (BW) per unit time
(e.g., mg/kgBW/day).
Adsorption.14 Surface retention of molecules, atoms, or ions by a solid or liquid, as opposed
to absorption, which is penetration of substances into the bulk of a solid or liquid.
Area Use Factor. The ratio of an organism's home range, breeding range, or
feeding/foraging range to the area of contamination of the site under investigation.
Glossary-1
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Assessment Endpoint. An explicit expression of the environmental value that is to be
protected.
Benthic Community.7 The community of organisms dwelling at the bottom of a pond, river,
lake, or ocean.
Bioaccurnulation.^ General term describing a process by which chemicals are taken up by an
organism either directly from exposure to a contaminated medium or by consumption of food
containing the chemical.
Bioccumulation Factor (BAF).3 The ratio of the concentration of a contaminant in an
organism to the concentration in the ambient environment at steady state, where the organism
can take in the contaminant through ingestion with its food as well as through direct contact.
Bipassay. Test used to evaluate the relative potency of a chemical by comparing its effect
on living organisms with the effect of a standard preparation on the same type of organism.
Bioassay and toxicity tests are not the same—see toxicity test. Bioassays often are run on a
series of dilutions of whole effluents.
Bioassessment. A general term referring to environmental evaluations involving living
organisms; can include bioassays, community analyses, etc.
Bioavailability.4 The degree to which a material in environmental media can be assimilated
by an organism.
Bioconcentration. A process by which there is a net accumulation of a chemical directly
from an exposure medium into an organism.
Biodegrade.15 Decompose into more elementary compounds by the action of living
organisms, usually referring to microorganisms such as bacteria.
Biomagnification.5 Result of the process of bioaccumulation and biotransfer by which tissue
concentrations of chemicals in organisms at one trophic level exceed tissue concentrations in
organisms at the next lower trophic level in a food chain.
Biomarker.21 Biochemical, physiological, and histological changes in organisms that can be
used to estimate either exposure to chemicals or the effects of exposure to chemicals.
Biomonitoring.5 Use of living organisms as "sensors" in environmental quality surveillance
to detect changes in environmental conditions that might threaten living organisms in the
environment.
Body Burden. The concentration or total amount of a substance in a living organism;
implies accumulation of a substance above background levels in exposed organisms.
Glossary-2
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Breeding Range. The area utilized by an organism during the reproductive phase of its life
cycle and during the time that young are reared.
Bulk Sediment.8 Field collected sediments used to conduct toxicity tests; can contain
multiple contaminants and/or unknown concentrations of contaminants.
Characterization of Ecological Effects.6 A portion of the analysis phase of ecological risk
assessment that evaluates the ability of a stressor to cause adverse effects under a particular
set of circumstances.
Characterization of Exposure.6 A portion of the analysis phase of ecological risk
assessment that evaluates the interaction of the stressor with one or more ecological
components. Exposure can be expressed as co-occurrence, or contact depending on the
stressor and ecological component involved.
Chemicals of Potential Concern.2 Chemicals that are potentially site-related and whose data
are of sufficient quality for use in a quantitative risk assessment.
Chronic.5 Involving a stimulus that is lingering or continues for a long time; often signifies
periods from several weeks to years, depending on the reproductive life cycle of the species.
Can be used to define either the exposure or the response to an exposure (effect). Chronic
exposures typically induce a biological response of relatively slow progress and long duration.
Chronic Response. The response of (or effect on) an organism to a chemical that is not
immediately or directly lethal to the organism.
Chronic Tests.9 A toxicity test used to study the effects of continuous, long-term exposure
of a chemical or other potentially toxic material on an organism.
Community.6 An assemblage of populations of different species within a specified location
and time.
Complexation.14 Formation of a group of compounds in which a part of the molecular
bonding between compounds is of the coordinate type.
Concentration. The relative amount of a substance in an environmental medium, expressed
by relative mass (e.g., mg/kg), volume (ml/L), or number of units (e.g., pans per million).
Concentration-Response Curve.5 A curve describing the relationship between exposure
concentration and percent of the test population responding.
Conceptual Model.6 Describes a series of working hypotheses of how the stressor might
affect ecological components. Describes ecosystem or ecosystem components potentially at
Glossary-3
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risk, and the relationships between measurement and assessment endpoints and exposure
scenarios.
Contaminant of (Ecological) Concern. A substance detected at a hazardous waste site that
has the potential to affect ecological receptors adversely due to its concentration, distribution,
and mode of toxicity.
Control.5 A treatment in a toxicity test that duplicates all the conditions of the exposure
treatments but contains no test material. The control is used to determine the response rate
expected in the test organisms in the absence of the test material.
Coordinate Bond.14 A chemical bond between two atoms in which a shared pair of
electrons forms the bond and the pair of electrons has been supplied by one of the two atoms.
Also known as a coordinate valence.
Correlation.10 An estimate of the degree to which two sets of variables vary together, with
no distinction between dependent and independent variables.
Critical Exposure Pathway. An exposure pathway which either provides the highest
exposure levels or is the primary pathway of exposure to an identified receptor of concern.
Degradation.14 Conversion of an organic compound to one containing a smaller number of
carbon atoms.
Deposition.14 The lying, placing, or throwing down of any material.
Depuration. A process that results in elimination of toxic substances from an organism.
Depuration Rate. The rate at which a substance is depurated from an organism.
Dietary Accumulation.9 The net accumulation of a substance by an organism as a result of
ingestion in the diet.
Direct Effect (toxin).6 An effect where the stressor itself acts directly on the ecological
component of interest, not through other components of the ecosystem.
Dose.11 A measure of exposure. Examples include (1) the amount of a chemical ingested,
(2) the amount of a chemical absorbed, and (3) the product of ambient exposure concentration
and the duration of exposure.
Dose-Response Curve.5 Similar to concentration-response curve except that the dose (i.e. the
quantity) of the chemical administered to the organism is known. The curve is plotted as
Dose versus Response.
Glossary-4
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Duplicate.8 A sample taken from and representative of the same population as another
sample. Both samples are carried through the steps of sampling, storage, and analysis in an
identical manner.
Ecological Component.6 Any part of an ecosystem, including individuals, populations,
communities, and the ecosystem itself.
Ecological Risk Assessment.6 The process that evaluates the likelihood that adverse
ecological effects may occur or are occurring as a result of exposure to one or more stressors.
Ecosystem. The biotic community and abiotic environment within a specified location and
time, including the chemical, physical, and biological relationships among the biotic and
abiotic components.
Ecotoxicity.11 The study of toxic effects on nonhuman organisms, populations, or
communities.
Estimated or Expected Environmental Concentration.5 The concentration of a material
estimated as being likely to occur in environmental media to which organisms are exposed.
Exposure. Co-occurrence of or contact between a stressor and an ecological component.
The contact reaction between a chemical and a biological system, or organism.
Exposure Assessment.2 The determination or estimation (qualitative or quantitative) of the
magnitude, frequency, duration, and route of exposure.
Exposure Pathway.2 The course a chemical or physical agent takes from a source to an
exposed organism. Each exposure pathway incudes a source or release from a source, an
exposure point, and an exposure route. If the exposure point differs from the source,
transport/exposure media (i.e., air, water) also are included.
Exposure Pathway Model. A model in which potential pathways of exposure are identified
for the selected receptor species.
Exposure Point.2 A location of potential contact between an organism and a chemical or
physical agent.
Exposure Point Concentration. The concentration of a contaminant occurring at an
exposure point.
Exposure Profile.6 The product of characterizing exposure in the analysis phase of
ecological risk assessment. The exposure profile summarizes the magnitude and spatial and
temporal patterns of exposure for the scenarios described in the conceptual model.
Glossary-5
-------
Exposure Route.2 The way a chemical or physical agent comes in contact with an organism
(i.e., by ingestion, inhalation, or dermal contact).
Exposure Scenario.6 A set of assumptions concerning how an exposure takes place,
including assumptions about the exposure setting, stressor characteristics, and activities of an
organism that can lead to exposure.
False Negative. The conclusion that an event (e.g., response to a chemical) is negative when
it is in fact positive (see Appendix D).
False Positive. The conclusion that an event is positive when it is in fact negative (see
Appendix D).
Fate. Disposition of a material in various environmental compartments (e.g. soil or
sediment, water, air, biota) as a result of transport, transformation, and degradation.
Food-Chain Transfer. A process by which substances in the tissues of lower-trophic-level
organisms are transferred to the higher-trophic-level organisms that feed on them.
Forage (feeding) Area. The area utilized by an organism for hunting or gathering food.
Habitat.1 Place where a plant or animal lives, often characterized by a dominant plant form
and physical characteristics.
Hazard. The likelihood that a substance will cause an injury or adverse effect under
specified conditions.
Hazard Identification.2 The process of determining whether exposure to a stressor can
cause an increase in the incidence of a particular adverse effect, and whether an adverse
effect is likely to occur.
Hazard Index.3 The sum of more than one hazard quotient for multiple substances and/or
multiple exposure pathways. The HI is calculated separately for chronic, subchronic, and
shorter-duration exposures.
\
Hazard Quotient.2 The ratio of an exposure level to a substance to a toxicity value selected
for the risk assessment for that substance (e.g., LOAEL or NOAEL).
Home Range.12 The area to which an animal confines its activities.
Hydrophilic.22 Denoting the property of attracting or associating with water molecules;
characteristic of polar or charged molecules.
Glossary-6
-------
Hydrophobic.12 With regard to a molecule or side group, tending to dissolve readily in
organic solvents, but not in water, resisting wetting, not containing polar groups or sub-
groups.
Hypothesis.12 A proposition set forth as an explanation for a specified phenomenon or group
of phenomena.
Indirect Effect.6 An effect where the stressor acts on supporting components of the
ecosystem, which in turn have an effect on the ecological component of interest.
Ingestion Rate. The rate at which an organism consumes food, water, or other materials
(e.g., soil, sediment). Ingestion rate usually is expressed in terms of unit of mass or volume
per unit of time (e.g., kg/day, L/day).
lonization. The process by which a neutral atom loses or gains electrons, thereby acquiring
a net charge and becoming an ion.
Lethal.5 Causing death by direct action.
Lipid.13 One of a variety of organic substances that are insoluble in polar solvents, such as
water, but that dissolve readily in non-polar organic solvents. Includes fats, oils, waxes,
steroids, phospholipids, and carotenes.
Lowest-Observable-Adverse-Effect Level (LOAEL). The lowest level of a stressor
evaluated in a toxicity test or biological field survey that has a statistically significant adverse
effect on the exposed organisms compared with unexposed organisms in a control or
reference site.
Matrix.14 The substance in which an analyte is embedded or contained; the properties of a
matrix depend on its constituents and form.
Measurement Endpoint.6 A measurable ecological characteristic that is related to the valued
characteristic chosen as the assessment endpoint. Measurement endpoints often are expressed
as the statistical or arithmetic summaries of the observations that make up the measurement.
As used in this guidance document, measurement endpoints can include measures of effect
and measures of exposure, which is a departure from U.S. EPA's (1992a) definition which
includes only measures of effect.
Media.15 Specific environmental compartments—air, water, soil—which are the subject of
regulatory concern and activities.
Median Effective Concentration (EC50).5 The concentration of a substance to which test
organisms are exposed that is estimated to be effective in producing some sublethal response
in 50 percent of the test population. The EC50 usually is expressed as a time-dependent value
Glossary-7
-------
(e.g., 24-hour EC50). The sublethal response elicited from the test organisms as a result of
exposure must be clearly defined.
Median Lethal Concentration (LC50).5 A statistically or graphically estimated
concentration that is expected to be lethal to 50 percent of a group of organisms under
specified conditions.
Metric. Relating to measurement; a type of measurement—for example a measurement of
one of various components of community structure (e.g., species richness, % similarity).
Mortality. Death rate or proportion of deaths in a population.
No-Observed-Adverse-Effect Level (NOAEL).5 The highest level of a stressor evaluated in
a toxicity test or biological field survey that causes no statistically significant difference in
effect compared with the controls or a reference site.
Nonparametric.17 Statistical methods that make no assumptions regarding the distribution of
the data.
Parameter.18 Constants applied to a model that are obtained by theoretical calculation or
measurements taken at another time and/or place, and are assumed to be appropriate for the
place and time being studied.
Parametric. Statistical methods used when the distribution of the data is known.
Population. An aggregate of individuals of a species within a specified location in space
and time.
Power.10 The power of a statistical test indicates the probability of rejecting the null
hypothesis when it should be rejected (i.e., the null hypothesis is false). Can be considered
the sensitivity of a statistical test. (See also Appendix D.)
Precipitation.14 In analytic chemistry, the process of producing a separable solid phase
within a liquid medium.
Precision.19 A measure of the closeness of agreement among individual measurements.
Reference Site.11 A relatively uncontaminated site used for comparison to contaminated sites
in environmental monitoring studies, often incorrectly referred to as a control.
Regression Analysis.10 Analysis of the functional relationship between two variables; the
independent variable is described on the X axis and the dependent variable is described on the
Y axis (i.e, the change in Y is a function of a change in X).
Glossary-8
-------
Replicate. Duplicate analysis of an individual sample. Replicate analyses are used for
quality control.
Representative Samples.18 Serving as a typical or characteristic sample; should provide
analytical results that correspond with actual environmental quality or the condition
experienced by the contaminant receptor.
Risk. The expected frequency or probability of undesirable effects resulting from exposure
to known or expected stressors.
Risk Characterization.6 A phase of ecological risk assessment that integrates the results of
the exposure and ecological effects analyses to evaluate the likelihood of adverse ecological
effects associated with exposure to the stressor. The ecological significance of the adverse
effects is discussed, including consideration of the types and magnitudes of the effects, their
spatial and temporal patterns, and the likelihood of recovery.
Sample.14 Fraction of a material tested or analyzed; a selection or collection from a larger
collection.
Scientific/Management Decision Point (SMDP). A point during the risk assessment process
when the risk assessor communicates results of the assessment at that stage to a risk manager.
At this point the risk manager determines whether the information is sufficient to arrive at a
decision regarding risk management strategies and/or the need for additional information to
characterize risk.
Sediment.20 Paniculate material lying below water.
Sensitivity. In relation to toxic substances, organisms that are more sensitive exhibit adverse
(toxic) effects at lower exposure levels than organisms that are less sensitive.
Sensitive Life Stage. The life stage (i.e., juvenile, adult, etc.) that exhibits the highest degree
of sensitivity (i.e., effects are evident at a lower exposure concentration) to a contaminant in
toxicity tests.
Species.13 A group of organisms that actually or potentially interbreed and are reproductively
isolated from all other such groups; a taxonomic grouping of morphologically similar
individuals; the category below genus.
Statistic.10 A computed or estimated statistical quantity such as the mean, the standard
deviation, or the correlation coefficient.
Stressor.6 Any physical, chemical, or biological entity that can induce an adverse response.
Glossary-9
-------
Sublethal.5 Below the concentration that directly causes death. Exposure to sublethal
concentrations of a substance can produce less obvious effects on behavior, biochemical
and/or physiological functions, and the structure of cells and tissues in organisms.
Threshold Concentration.5 A concentration above which some effect (or response) will be
produced and below which it will not.
Toxic Mechanism of Action.23 The mechanism by which chemicals produce their toxic
effects, i.e., the mechanism by which a chemical alters normal cellular biochemistry and
physiology. Mechanisms can include; interference with normal receptor-ligand interactions,
interference with membranae functions, interference with cellular energy production, and
binding to biomolecules.
Toxicity Assessment. Review of literature, results in toxicity tests, and data from field
surveys regarding the toxicity of any given material to an appropriate receptor.
Toxicity Test.5 The means by which the toxicity of a chemical or other test material is
determined. A toxicity test is used to measure the degree of response produced by exposure
to a specific level of stimulus (or concentration of chemical) compared with an unexposed
control.
Toxicity Value.2 A numerical expression of a substance's exposure-response relationship that
is used in risk assessments.'
Toxicant. A poisonous substance.
Trophic Level.6 A functional classification of taxa within a community that is based on
feeding relationships (e.g., aquatic and terrestrial plants make up the first trophic level, and
herbivores make up the second).
Type I Error.10 Rejection of a true null hypothesis (see also Appendix D).
Type II Error.10 Acceptance of a false null hypothesis (see also Appendix D).
Uptake.5 A process by which materials are transferred into or onto an organism.
Uncertainty.] ] Imperfect knowledge concerning the present or future state of the system
under consideration; a component of risk resulting from imperfect knowledge of the degree of
hazard or of its spatial and temporal distribution.
Volatilization.14 The conversion of a chemical substance from a liquid or solid state to a
i
gaseous vapor state.
Glossary-10
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Xenobiotic.6 A chemical or other stressor that does not occur naturally in the environment.
Xenobiotics occur as a result of anthropogenic activities such as the application of pesticides
and the discharge of industrial chemicals to air, land, or water.
ENDNOTES
1 Krebs 1978, 2 U.S. EPA 1989, 3 Calow 1993, 4 Freedman 1989, 5 Rand and Petrocelli
1985, 6 U.S. EPA 1992a, 7 Ricklefs 1990, 8 U.S. EPA 1992b, 9 ASTM 1993a, 10 Sokal and
Rohlf 1981, n Suter 1993, 12 Wallace et al. 1981, 13 Curtis 1983, 14 Parker 1994, 15 Sullivan
1993, 16 U.S. EPA 1990, 17 Zar 1984, 18 Keith 1988, 19 Gilbert 1987, 20 ASTM 1993b,
21 Huggett et al. 1992, ^ Stedman 1995, 23 Amdur et al. 1991.
GLOSSARY REFERENCES
Amdur, M.O.; Doull J.; Klaassen, C.D. 1991. Casarett and Doull's Toxicology. Fourth
Edition. New York, NY: McGraw-Hill. .
American Society for Testing and Materials (ASTM). 1993a. ASTM Standard E 943.
Standard terminology relating to biological effects and environmental fate.
American Society for Testing and Materials (ASTM). 1993b. ASTM Standard E 1525.
Standard guide for designing biological tests with sediments.
Calow, P. (ed.). 1993. Handbook of Ecotoxicology. Volume!. Boston, MA: Blackwell
Publishing.
Curtis, H. 1983. Biology. Fourth Edition. New York, NY: Worth.
Freedman, B. 1989. Environmental Ecology. The Impacts of Pollution and Other Stresses
on Ecosystem Structure and Function. New York, NY: Academic Press.
Gilbert, R.O. 1987. Statistical Methods for Environmental Pollution Monitoring. New York,
NY: Reinhold.
Keith, L.H. (ed.). 1988. Principles of Environmental Sampling. American Chemical Society.
Krebs, C.J. 1978. Ecology: The experimental analysis of distribution and abundance.
Second edition. New York, NY: Harper & Row.
Huggett, R.J.; Kimerle, R.A.; Nehrle, P.M. Jr.; Bergman, H.L. (eds.). 1992. Biomarkers:
Biochemical, Physiological, and Histological Markers of Anthropogenic Stress. A
Special Publication of SET AC. Chelsea, MI: Lewis Publishers.
Glossary-11
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Parker, S.P. (ed.). 1994. Dictionary of Scientific and Technical Terms. Fifth Edition. New
York, NY: McGraw-Hill.
Rand, G.M.; Petrocelli, S.R. 1985. Fundamentals of Aquatic Toxicology. Methods and
Applications. New York, NY: McGraw Hill.
Ricklefs, R.E. 1990. Ecology. Second Edition. New York, NY: W.H. Freeman.
Sokal, R.R.; Rohlf, FJ. 1981. Biometry. Second Edition. New York, NY: W.H. Freeman.
Stedman, T.L. 1995. Stedman's Medical Dictionary. 26th Edition. Baltimore, MD:
Williams and Wilkins.
Sullivan, T.F.P. 1993. Environmental Regulatory Glossary. Government Institutes, Inc.
Suter, G.W. n. 1993. Ecological Risk Assessment. Ann Arbor, ML Lewis.
U. S. Environmental Protection Agency (U.S. EPA). 1989. Risk Assessment Guidance for
Superfund: Volume 1 - Human Health. Washington, DC: Office of Emergency and
Remedial Response; EPA/540/1-89/002.
U. S. Environmental Protection Agency (U.S. EPA). 1990. Macroinvertebrate Field and
Laboratory Methods for Evaluating the Biological Integrity of Surface Waters.
Washington, DC: Office of Water; EPA/600/4-90/030.
U. S. Environmental Protection Agency (U.S. EPA). 1992a. Framework for Ecological Risk
Assessment. Washington, DC: Risk Assessment Forum; EPA/630/R-02/011.
U.S. Environmental Protection Agency (U.S. EPA). 1992b. Sediment Classification Methods
Compendium. Washington, DC: Office of Water; EPA/823/R-092/006.
Wallace, R.A.; King, J.L.; Sanders, G.P. 1981. Biology. The Science of Life. Second
Edition. IL: Scott, Foresman & Co.
Zar, J.H. 1984. Biostatistical Analysis. Princeton, NJ: Prentice-Hall.
Glossary-12
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APPENDIX A
EXAMPLE ECOLOGICAL RISK ASSESSMENTS
FOR HYPOTHETICAL SITES
-------
INTRODUCTION
Appendix A provides examples of Steps 1 through 5 of the ecological risk assessment
process for three hypothetical sites:
(1) A former municipal landfill from which copper is leaching into a large pond
down-gradient of the site (the copper site);
(2) A former chemical production facility that spilled DDT, which has been
transported into a nearby stream by surface water runoff (the DDT site); and
(3) A former waste-oil recycling facility that disposed of PCBs in a lagoon from
which extensive soil contamination has resulted (the PCB site).
These examples are intended to illustrate key points in Steps 1 through 5 of the ecological
risk assessment process. No actual site is the basis for the examples.
The examples stop with Step 5 because the remaining steps (6 through 8) of the
ecological risk assessment process and the risk management decisions depend on site-specific
data collected during a site investigation. We have not attempted to develop hypothetical data
for analysis or the full range of information that a site risk manager would consider when
evaluating remedial options.
-------
EXAMPLE 1: COPPER SITE
STEP 1: SCREENING-LEVEL PROBLEM FORMULATION AND ECOLOGICAL
EFFECTS EVALUATION
Site history. This is a former municipal landfill located in an upland area of the
mid-Atlantic plain. Residential, commercial, and industrial refuse was disposed of at this site
in the 1960s and 1970s. Large amounts of copper wire also were disposed at this site over
several years. Currently, minimal cover has been placed over the fill and planted with
grasses. Terrestrial ecosystems in the vicinity of the landfill include upland forest and
successional fields. Nearby land uses include agriculture and residential and commercial uses.
The landfill cover has deteriorated in several locations. Leachate seeps have been noted on
the slope of the landfill, and several seeps discharge to a five-acre pond down-gradient of the
site.
Site visit. A preliminary site visit was conducted and the ecological checklist was
completed. The checklist indicated that the pond has an organic substrate; emergent
vegetation, including cattail and rushes, occurs along the shore near the leachate seeps; and
the pond reaches a depth of five feet toward the middle. Fathead minnows, carp, and several
species of sunfish were observed, and the benthic macroinvertebrate community appeared to
be diverse. The pond water was clear, indicating an absence of phytoplankton. The pond
appears to function as a valuable habitat for fish and other wildlife using this area.
Preliminary sampling indicated elevated copper levels in the seep as well as elevated base
cations, total organic carbon (TOC), and depressed pH levels (pH 5.7).
Problem formulation. Copper is leaching from the landfill into the pond from a
seep area. EPA's ambient water quality criteria document for copper (U.S. EPA, 1985)
indicates that it can cause toxic effects in aquatic plants, aquatic invertebrates, and young fish
at relatively low water concentrations. Thus, the seep might threaten the ability of the pond
to suppon macroinvertebrate and fish communities and the wildlife that feed on them.
Terrestrial ecosystems do not need to be evaluated because the overland flow of the seeps is
limited to short gullies, a few inches wide. Thus, the area of concern has been identified as
the five-acre pond and the associated leachate seeps. Copper in surface water and sediments
of the pond might be of ecological concern.
Ecological effects evaluation. Copper is toxic to both aquatic plants and aquatic
animals. Therefore, aquatic toxicity-based data will be used to screen for ecological risk in
the preliminary risk calculation. The screening ecotoxicity value selected for water-column
exposure is the U.S. EPA chronic ambient water quality criterion (12 ug/L at a water, hardness
of 100 mg/L as CaCO3). A screening ecotoxicity value for copper in sediments was
identified as 34 mg/kg (U.S. EPA, 1996).
A-l
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STEP 2: SCREENING-LEVEL EXPOSURE ESTIMATE AND RISK CALCULATION
Exposure estimate. Preliminary sampling data indicate that the leachate contains
53 ug/L copper as well as elevated base cations, elevated TOC, and depressed pH (pH 5.7). '
Sediment concentrations range from 300 mg/kg to below detection (2 mg/kg), decreasing with
distance from the leachate seeps.
i
Risk calculation. The copper concentration in the seep water (53 Mg/L) exceeds the
chronic water quality criterion for copper (12 Mg/L). The maximum sediment copper
concentration of 300 mg/kg exceeds the screening ecotoxicity value for copper in sediments
(34 mg/kg). Therefore, the screening-level hazard quotients for both sediment and water
exceed one. The decision at the Scientific/Management Decision Point (SMDP) is to continue
the ecological risk assessment
Similar screening for the levels of base cations generated hazard quotients below one
in the seep water. .Although TOC and pH are not regulated under CERCLA, the possibility
that those parameters might affect the biota of the pond should be kept in mind if surveys of
the pond biota are conducted. Sediment concentrations of chemicals other than copper
generated hazard quotients (HQs) of less than one at the maximum concentrations found.
STEP 3: BASELINE RISK ASSESSMENT PROBLEM FORMULATION
Based on the screening-level risk assessment, copper is known to be the only
contaminant of ecological concern at the site.
Ecotoxicity literature review. A review of the literature on the ecotoxicity of
copper to aquatic biota was conducted and revealed several types of information. Young
aquatic organisms are more sensitive to copper than adults (Demayo et al., 1982; Kaplan and
Yoh, 1961; Hubschman, 1965). Fish larvae usually are more sensitive than embryos (McKim
et al., 1978; Weis and Weis, 1991), and fish become less sensitive to copper as body weight
increases (Demayo et al., 1982). Although the exact mechanism of toxicity to fish is
unknown, a loss of osmotic control has been noted in some studies (Demayo et al. 1982;
Cheng and Sullivan, 1977).
Flowthrough toxicity studies in which copper concentrations were measured revealed
LC50 values ranging from 75 to 790 ug/L for fathead minnows and 63 to 800 ug/L for
common carp (U.S. EPA, 1985). Coldwater fish species, such as rainbow trout, can be more
sensitive, and species like pumpkinseeds (a sunfish) and bluegills are less sensitive (U.S.
EPA, 1985). Although fish fry usually are the most sensitive life stage, this is not always the
case; Pickering et al. (1977) determined an LC50 of 460 ug/L to 6-month-old juveniles and an
LC50 of 490 pg/L to 6-week-old fry for fathead minnows. A copper concentration in water
of 37 ug/1 has been shown to cause a significant reduction in fish egg production (Pickering
et al., 1977).
A-2
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Elevated levels of copper in sediments have been associated with changes in benthic
community structure, notably reduced numbers of species (Winner et al., 1975; Kraft and
Sypniewski, 1981). Studies also have been conducted with adult Hyalella azteca (an
amphipod) exposed to copper in sediments. One of these studies indicated an LC50 of 1,078
mg/kg in the sediment (Cairns et al., 1984); however, a no-observed-adverse-effect level
(NOAEL) for copper in sediments was not identified for an early life stage of a benthic
invertebrate.
A literature review of the ecotoxicity of copper to aquatic plants, both algae and
vascular plants, did not reveal information on the toxic mechanism by which copper affects
plants. The review did indicate that exposure of plants to high copper levels inhibits
photosynthesis and growth (U.S. EPA, 1985), and cell separation after cell division (Hatch,
1978). Several studies conducted using Selenastrum capricornutum indicated that
concentrations at 300 ug/L kill algae after 7 days, and a value of 90 ug/1 causes complete
growth inhibition after 7 days (Bartlett et al., 1974).
The literature indicates that copper does not biomagnify in food chains and does not
bioaccumulate in most animals because it is a biologically regulated essential element.
Accumulation in phytoplankton and filter-feeding mollusks, however, does occur. The
toxicity of copper in water is influenced by water hardness, alkalinity, and pH (U.S. EPA,
1985).
Exposure pathways. A flow diagram was developed to depict the environmental
pathways that could result in impacts of copper to the pond's biota (see Exhibit A-l). Direct
exposure to copper in the pond water and sediments could cause acute or chronic toxicity in
early life stages of fish and/or benthic invertebrates, and in aquatic plants. Risks to filter-
feeding mollusks and phytoplankton as well as animals that feed on them are not considered
because the mollusks and phytoplankton are unlikely to occur in significant quantities in the
pond. The exposure pathways that will be evaluated, therefore, are direct contact with
contaminated sediments and water.
Assessment endpoints and conceptual model. Based on the screening-level
risk assessment, the ecotoxicity literature review, and the complete exposure pathways,
development of a conceptual model for the site is initiated. Copper can be acutely or
chronically toxic to organisms in an aquatic community through direct exposure of the
organisms to copper in the water and sediments. Threats of copper to higher trophic level
organisms are unlikely to exceed threats to organisms at the base of the food chain, because
copper is an essential nutrient which is effectively regulated by most organisms if the
exposure is below toxic levels. Fish fry in particular can be very sensitive to copper in water.
Based on these receptors and the potential for both acute and chronic toxicity, an
appropriate general assessment endpoint for the ecosystem would be the maintenance of the
community composition of the pond. A more operational definition of the assessment
endpoint would be the maintenance of pond community structure typical for the locality and
A-3
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for the physical attributes of the pond, with no loss of species or community alteration due to
copper toxicity.
Risk questions. One question is whether the concentrations of copper present in the
sediments and water over at least pan of the pond are toxic to aquatic plants or animals. A
further question is what concentration of copper in sediments represents a threshold for
adverse effects. That level could be used as a preliminary cleanup goal.
STEP 4: MEASUREMENT ENDPOINTS AND STUDY DESIGN
To answer the hypothesis identified in Step 3, three lines of evidence were considered
when selecting measurement endpoints: (1) whether the ambient copper levels are higher
than levels known to be directly toxic to aquatic organisms likely or known to be present in
the pond; (2) whether water and sediments taken from the pond are more toxic to aquatic
organisms than water and sediments from a reference pond; and (3) whether the aquatic
community structure in the site pond is simplified relative to a reference pond.
Measurement endpoints. Since the identified assessment endpoint is maintaining
a typical pond community structure, the possibility of directly measuring the condition of the
plant, fish, and macroinvertebrate communities in the pond was considered. Consultation with
experts on benthic macroinvertebrates suggested that standard measures of the pond benthic
invertebrate community probably would be insensitive measures of existing effects at this
particular site because of the high spatial variation in benthic communities within and among
ponds of this size. Measuring the fish community also would be unsuitable, due to the
limited size of the pond and low diversity of fish species anticipated. Since copper is not
expected to bioaccumulate or biomagnify in this pond, direct toxicity testing was selected as
appropriate. Because early life stages tend to be more sensitive to the toxic effects of copper
than older life stages, chronic toxicity would be measured on early life stages. For animals,
toxicity is defined as a statistically significant decrease in survival or juvenile growth rates
(measurement endpoints) of a test group exposed to water or sediments from the site
compared with a test group exposed to water or sediments from a reference site. For plants,
toxicity is defined as a statistically significant decrease in growth rate (measurement endpoint)
with the same comparison.
One toxicity test selected is a 10-day (i.e., chronic) solid-phase sediment toxicity test
using an early life stage of Hyalella azteca. The measures of effects for the test are mortality
rates and growth rates (measured as length and weight increases). Two water-column toxicity
tests will be used: (1) a 7-day test using the alga Selenasirum capricornutum (growth test)
and (2) a 7-day larval fish test using Pimephales promelas (mortality and growth endpoints).
The H. azteca and P. promelas toxicity tests will be used to determine the effects of copper
on early life stages of invertebrates and fish in sediment and the water column, respectively.
The test on S. capricornutum will be used to determine the phytotoxicity of copper in the
water column.
A-4
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EXHIBIT A-1
Conceptual Model for the Copper Site
MEASUREMENT ENDPOINT
(Sediment loxicity to Hyatella azieca)
PRIMARY SOURCE
(Landfill)
h
SECONDARY
SOURCE
(Groundwater seep)
TERTIARY SOURCE
(Sediment, exposure point
for aquatic receptors)
TERTIARY SOURCE
(Surface water, exposure
point for aquatic receptors)
ASSESSMENT
ENDPOINT
AQUATIC RECEPTOR
ASSESSMENT
ENDPOINT
AQUATIC RECEPTOR
MEASUREMENT ENDPOINT
(Surface water toxicity to Selenastrum
capricornatum and Pimephalespromelas)
-------
Study design. To answer the questions stated in the problem formulation step, the
study design specified in the following. The water column tests will be run on 100 percent
seep water, 100 percent pond water near the seep, 100 percent reference-site water, and the
laboratory control. U.S. EPA test protocols will be followed. Five sediment samples will be
collected from the pond bottom at intervals along the observed concentration gradient, from a
copper concentration of 300 mg/kg at the leachate seeps down to approximately 5 mg/kg near
the other end of the pond. The sediment sampling locations will transect the pond at
equidistant locations and include the point of maximum pond depth. All sediment samples
will be split so that copper concentrations can be measured in sediments from each sampling
location. A reference sediment will be coUected and a laboratory control will be run. Test .
organisms will not be fed during the test; sediments will be sieved to remove native
organisms and debris. Laboratory procedures will follow established protocols and will be
documented and reviewed prior to initiation of the test. For the water-column test, statistical
comparisons will be made between responses to each of the two pond samples and the
reference site, as well as the laboratory control. Statistical comparisons also will be made of
responses to sediments taken from each sampling location and responses to the reference
sediment sample.
Because leachate seeps can be intermittent (depending on rainfall), the study design
specifies that a pre-sampling visit is required to confirm that the seep is flowing and can be
sampled. The study design also specifies that both sediments and water will be sampled at
the same time at each sampling location.
As the work plan (WP) and sampling and analysis plan (SAP) were finished, the
ecological risk assessor and the risk manager agreed on the site conceptual model, assessment
endpoints, and study design (SMDP).
STEPS: FIELD VERIFICATION OF STUDY DESIGN
A site assessment was conducted two days prior to the scheduled initiation of the site
investigation to confirm that the seep was active. It was determined that the seep was active
and that the site investigation could be initiated.
REFERENCES
Bartlett, L.; Rabe, F.W.; Funk, W.H. 1974. Effects of copper, zinc, and cadmium on
Selenastrum capricornutum. Water Res. 8: 179-185.
Cairns, M.A.; Nebeker, A.V.; Gakstatter, J.H.; Griffis, W.L. 1984. Toxicity of copper-spiked
sediments to freshwater invertebrates. Environ. Toxicol. Chem. 3: 345-445.
Cheng, T.C.; Sullivan, J.T. 1977. Alterations in the osmoregulation of the pulmonate
gastropod Biomphalaria glabrata due to copper. J. Invert. Path. 28: 101.
A-6
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Demayo, A., et al. 1982. Effects of copper on humans, laboratory and farm animals,
terrestrial plants, and aquatic life. CRC Crit. Rev. Environ. Control. 12: 183.
Hatch, R.C. 1978. Poisons causing respiratory insufficiency. In: L.M. Jones, N.H, Booth
and L.E. McDonald (eds.), Veterinary Pharmacology and Therapeutics. Iowa State
University, IA: Ames Press.
Hubschman, J.H. 1965. Effects of copper on the crayfish Orconectes rusticus (Girard). I.
Acute toxicity. Crustaceana 12: 33-42.
Kaplan, H.M.; Yoh, L. 1961. Toxicity of copper to frogs. Herpetologia 17: 131-135.
Kraft, KJ.; Sypniewski, R.H. 1981. Effect of sediment copper on the distribution of benthic
macroinvertebrates in the Keweenaw Waterway. J. Great Lakes Res. 7: 258-263.
McKim, J.M.; Eaton, J.G.; Holcombe, G.W. 1978. Metal toxicity to embryos and larvae of
eight species of freshwater fish. II. Copper. Bull. Environ. Contain. Toxicol. 19:
608-616.
Pickering, Q.; Brungs, W.; Gast, M. 1977. Effect of exposure time and copper concentration
of fathead minnows, Pimephales promelas (Rafinesque). Aquatic Toxicol. 12: 107.
U.S. Environmental Protection Agency (U.S. EPA). 1996. Ecotox Thresholds. ECO Update,
Intermittent Bulletin, Volume 3, Number 2. Washington, DC: Office of Emergency
and Remedial Response, Hazardous Site Evaluation Division; Publication 9345.0-
12FSI; EPA/540/F-95/038; NTIS PB95-963324.
U.S. Environmental Protection Agency (U.S. EPA). 1985. Ambient Water Quality Criteria
for Copper. Washington, DC: Office of Water; EPA/440/5-84/031.
Weis, P.; Weis, J.S. 1991. The developmental toxicity of metals and metalloids in fish. In:
Newman, M.C.; Mclntosh, A.W. (eds.), Metal Ecotoxicology: Concepts and
Applications. Boca Raton, FL: CRC Press, Inc., Lewis Publishers.
Winner, R.W.; Kelling, T.; Yeager, R.; et al. 1975. Response of a macroinvertebrate fauna
to a copper gradient in an experimentally-polluted stream. Verb. Int. Ver. Limnol. 19:
2121-2127. ,
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EXAMPLE 2: DDT SITE
STEP 1: SCREENING-LEVEL PROBLEM FORMULATION AND ECOLOGICAL
EFFECTS EVALUATION
Site history. This is the site of a former chemical production facility located
adjacent to a stream. The facility manufactured and packaged dichlorodiphenyltrichloroethane
(DDT). Due to poor storage practices, several DDT spills have occurred.
Site visit A preliminary site visit was conducted and the ecological checklist was
completed. Information gathered indicates that surface water drainage from the site flows
through several drainage swales toward an unnamed creek. This creek is a second-order
stream containing riffle-run areas and small pools. The stream substrate is composed of sand
and gravel in the pools with some depositional areas in the backwaters and primarily cobble
in the riffles.
Problem formulation. Previous sampling efforts indicated the presence of DDT
and its metabolites in the stream's sediments over several miles at concentrations up to
230 mg/kg. A variety of wildlife, especially piscivorous birds, use this area for feeding.
Many species of minnow have been noted in this stream. DDT is well known for its
tendency to bioaccumulate and biomagnify in food chains, and available evidence indicates
that it can cause reproductive failure in birds due to eggshell thinning.
The risk assessment team and risk manager agreed that the assessment endpoint is
adverse effects on reproduction of high-trophic-level wildlife, particularly piscivorous birds.
Ecological effects evaluation. Because DDT is well studied, a dietary
concentration above which eggshell thinning might occur was identified in existing U.S. EPA
documents on the ecotoxicity of DDT. Moreover, a no-observed-adverse-effect-level
(NOAEL) for the ingestion route for birds also was identified.
STEP 2: SCREENING-LEVEL EXPOSURE ESTIMATE AND RISK CALCULATION
Exposure estimate. For the screening-level exposure estimate, maximum
concentrations of DDT identified in the sediments were used. To estimate the concentration
of DDT in forage fish, the maximum concentration in sediments was multiplied by the
highest DDT bioaccumulation factor relating forage fish tissue concentrations to sediment
concentrations reported in the literature. Moreover, it was assumed that the piscivorous birds
obtain 100 percent of their diet from the contaminated area.
Risk calculation. The predicted concentrations of DDT in forage fish were
compared with the dietary NOAEL for DDT in birds. This risk screen indicated that DDT
concentrations measured at this site might be high enough to cause adverse reproductive
A-8
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effects in birds. Thus, transfer of DDT from the sediments to the stream and biota are of
concern at this site.
STEP 3: BASELINE RISK ASSESSMENT PROBLEM FORMULATION
Based on the screening-level risk assessment, potential bioaccumulation of DDT in
aquatic food chains and effects of DDT on reproduction in piscivorous birds are known
concerns. During refinement of the problem, the potential for additional ecological effects of
DDT was examined.
Ecotoxicity literature review. In freshwater systems, DDT can have direct effects
on animals, particularly aquatic insects. A literature review of the aquatic toxicity of DDT
was conducted, and a NOAEL and LOAEL identified for the toxicity of DDT to aquatic
insects. Aquatic plants are not affected by DDT. Additional quantitative information on
effects of DDT on birds was reviewed, particularly to identify what level of eggshell thinning
is likely to reduce reproductive success. A number of studies have correlated DDT residues
measured in eggs of birds to increased eggshell thinning arid egg loss due to breakage.
Eggshell thinning of more than 20 percent appears to result in decreased hatching success due
to eggshell breakage (Anderson and Hickey, 1972; Dilworth et al., 1972). Information was
not available for any piscivorous species of bird. Lincer (1975) conducted a laboratory
feeding study using American kestrels. Females fed a diet of 6 mg/kg DDE1
(1.1 mg/kgBW-day) produced eggs with shells which were 25.5 percent thinner than archived
eggshells collected prior to widespread use of DDT. Based on this information, a LOAEL of
1.1 mg/kgBW-day was selected to evaluate the effects of DDT on piscivorous birds.
Exposure pathways, assessment endpoints, and conceptual model. Based
on knowledge of the fate and transport of DDT in aquatic systems and the ecotoxicity of
DDT to aquatic organisms and birds, a conceptual model was initiated. DDT buried in the
sediments can be released to the water column during resuspension and redistribution of the
sediments. Some diffusion of DDT to the water column from the sediment surface also will
occur. The benthic community would be an initial receptor for the DDT in sediments, which
could result in reduced benthic species abundance and DDT accumulation in species that
remain. Fish that feed on benthic organisms might be exposed to DDT both in the water
column and in their food. Piscivorous birds would be exposed to the DDT that has
accumulated in the fish, and could be exposed at levels sufficiently high to cause more than
20 percent eggshell thinning. Based on this information, two assessment endpoints were
identified: (1) maintaining stream community structure typical for the stream order and
location, and (2) protecting piscivorous birds from eggshell thinning that could result in
reduced reproductive success.
1 DDE is a degradation product of DDT; typically, field measures of DDT are reported as the sum of the
concentrations of DDT, DDE, and DDD (another degradation product).
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A flow diagram of the exposure pathways for DDT was added to the conceptual model
(Exhibit A-2). The diagram identifies the primary, secondary, and tertiary sources of DDT at
the site, as well as the primary, secondary, and tertiary types of receptors that could be -
exposed.
Risk questions. Two questions were developed: (1) has the stream community
been affected by the DDT, and (2) have food-chain accumulation and transfer of DDT
occurred to the extent that 20 percent or more eggshell thinning would be expected hi
piscivorous birds that use the area.
STEP 4: MEASUREMENT ENDPOINTS AND STUDY DESIGN
Measurement endpoints. For the assessment endpoint of protecting piscivorous
birds from eggshell thinning, the conceptual model indicated that DDT in sediments could
reach piscivorous birds through forage fish. Belted kingfishers are known to feed in the
stream. They also have the smallest home range of the piscivorous birds hi the area, which
means that more kingfishers can forage entirely from the contaminated stream area than can
other species of piscivorous birds. Thus, one can conclude that, if the risk assessment shows
no threat of eggshell thinning to the kingfisher, there should be minimal or no threat to other
piscivorous birds that might utilize the site. Eggshell thinning hi the belted kingfisher
therefore was selected as the measure of effect.
Data from the literature suggest that DDT can have a bioaccumulation factor in
surface water systems as high as six orders of magnitude (106); however, in most aquatic
ecosystems, the actual bioaccumulation of DDT from the environment is lower, often
substantially lower. Many factors influence the actual accumulation of DDT in the
environment. There is considerable debate over the parameters of any proposed theoretical
bioaccumulation model; therefore, it was decided to measure tissue residue levels in the
forage fish at the site instead of estimating the tissue residue levels in forage fish using a
bioaccumulation factor (BAF).
Existing information on the distribution of DDT in the stream indicates that a general
gradient of DDT concentrations exists in the sediments, and five locations could be identified
that corresponded to a range of DDT concentrations in sediments. Based on information
available on fish communities in streams.similar to the one in the site area, creek chub
(Semotilus atromaculaius) were selected to measure exposure levels for kingfishers. Creek
chub feed on benthic invertebrates, which are in direct contact with the contaminated
sediments. Adult creek chub average 10 inches and about 20 grams, allowing for analysis of
individual fish. Creek chub also have small home ranges during the spring and summer, and
thus it should be possible to relate DDT levels in the chub to DDT levels in the sediments.
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EXHIBIT A-2
Conceptual Model for the Stream DDT Site
MEASUREMENT ENDPOINT
(DDT concentration in fish
tissue, exposure point for
kingfishers)
SECONDARY
RECEPTOR
(Fish)
PRIMARY SOURCE
(Plant site)
SECONDARY
SOURCE
(Surface drainage)
ASSESSMENT
ENDPOINT
TERTIARY RECEPTOR
(Piscivorous bird)
TERTIARY SOURCE
(Stream sediments,
exposure point for fish and
macroinvertebrales)
PRIMARY RECEPTOR
(Benthic
macroinvertebrates,
exposure point for fish)
MEASUREMENT ENDPOINT
(Benthic macroinvertebrate
community structure)
-------
For the assessment endpoint of maintaining stream community structure, the selected
measurement endpoints were several metrics describing the abundance and trophic structure of
the stream benthic macroinvertebrate community.
Study design. The study design specified that creek chub would be collected at
several locations with known DDT concentrations in sediments. The fish would be analyzed
for body burdens of DDT, and the relationship between DDT levels in the sediments and in
the creek chub would be established. The fish DDT concentrations would be used to evaluate
the DDT threat to piscivorous birds feeding on the fish at each location. Using the DDT
concentrations measured in fish that correspond to a LOAEL and NOAEL for adverse effects
in birds, the corresponding sediment'contamination levels would be determined. Those
sediment DDT levels then could be used to derive a cleanup level that would reduce threats
of eggshell thinning to piscivorous birds.
The study design for measuring DDT residue levels in creek chub specified that
10 creek chub of the same size and sex would be collected at each location and that each
creek chub be at least 20 grams, so that individuals could be analyzed. In addition, at one
location, QA/QC requirements dictated that an additional 10 fish be collected. In this
example, it was necessary to verify in the field that sufficient numbers of creek chub of the '
specified size were present to meet the tissue sampling requirements. In addition, the stream
conditions needed to be evaluated to determine what fish sampling techniques would work
best at the targeted locations.
The study design and methods for benthic macroinvertebrate collection followed the
Rapid Bioassessment Protocol (RBP) manual for level three evaluation (U.S. EPA, 1989).
Benthic macroinvertebrate samples were co-located with sampling for fish tissue residue
levels so that one set of co-located water and sediment samples for analytic chemistry could
serve for comparison with both tissue analyses.
The study design also specified that the hazard quotient (HQ) method would be used
to evaluate the effects of DDT on the kingfisher during risk characterization. To determine
the HQ, the estimated daily dose of DDT consumed by the kingfishers is divided by a
LOAEL of 1.1 mg/kgBW-day for kestrels. To estimate the DDT dose to the kingfisher, the
DDT concentrations in the chub is multiplied by the fish ingestion rate for kingfishers and
divided by the body weight of kingfishers. This dose is adjusted by the area use factor. The
area use factor corresponds to the proportion of the diet of a kingfisher that would consist of
fish from the contaminated area. The area use factor is a function of the home range size of
kingfishers relative to the area of contamination. The adjusted dose is compared to the
LOAEL. A HQ of greater than one implies that impaired reproductive success in kingfishers
due to site contamination is likely, and an HQ of less than one implies impacts due to site
contaminants are unlikely (see text Section 2.3 for a description of HQs).
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STEP 5: FIELD VERIFICATION OF STUDY DESIGN
A field assessment was conducted and several small fish collection techniques were
used to determine which technique was the most effective for capturing creek chub at the site.
Collected chub were examined to determine the size range available and to determine if
individuals could be sexed.
Seine netting the areas targeted indicated that the creek chub might not be present in
sufficient numbers to provide the necessary biomass for chemical analyses. Based on these
findings, a contingency plan was agreed to (SMDP), which stated that both the creek chub
and the longnosed dace (Rhinichthys cataractae) would be collected. If the creek chub were
collected at all locations in sufficient numbers, those samples would be analyzed and the dace
would be released. If sufficient creek chub could not be collected but sufficient longnosed
dace could, the longnosed dace would be analyzed and the creek chub released. If neither
species could be collected at all locations in sufficient numbers, then a mix of the two species
would be used; however, for any given site only one species would be analyzed. In addition,
at one location, preferably one with high DDT levels in the sediment, sufficient numbers of
approximately 20 gram individuals of both species would be collected to allow comparison
(and calibration) of the accumulation between the two species. If necessary to meet the
analytic chemistry needs, similarly-sized individuals of both sexes of creek chub would be
pooled. Pooling two or more individuals would be necessary for the smaller dace. The risk
assessment team decided that the fish samples would be collected by electro-shocking. Field
notes for all samples would document the number of fish per sample pool, sex, weight,
length, presence of parasites or deformities, and other measures and might help to explain any
anomalous data.
REFERENCES
Anderson, D.W.; Hickey, JJ. 1972. Eggshell changes in certain North American birds. In:
Voos, K.H. (ed.), Proceedings: XV International Ornithological Congress. The
Hague, Netherlands; pp. 514-540.
Dilworth, T.G., Keith, J.A.; Pearce, P.A.; Reynolds, L.M. 1972. DDE and eggshell thickness
in New Brunswick woodcock. J. Wildl. Manage. 36: 1186-1193.
Lincer, J.L. 1975. DDE-induced eggshell thinning in the American kestrel; a comparison of
the field situation and laboratory results. J. App. Ecol. 12: 781-793.
U.S. Environmental Protection Agency (U.S. EPA). 1989. Rapid Bioassessment Protocols
for Use in Streams and Rivers: Benthic Macroinvertebrates and Fish. Washington,
DC: Office of Water (Plafkin, J.L., Barbour, M.T., Porter, K.D., Gross, S.K., and
Hughes, R.M., authors); EPA/440/4-89/001.
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EXAMPLE 3: PCB SITE
STEP 1: SCREENING-LEVEL PROBLEM FORMULATION AND ECOLOGICAL
EFFECTS EVALUATION
Site history. This is a former waste-oil recycling facility located in a remote area.
Oils contaminated with polychlorinated biphenyl compounds (PCBs) were disposed of in a
lagoon. The lagoon was not lined, and the soil is composed mostly of sand. Oils
contaminated with PCBs migrated through the soil and contaminated a wide area adjacent to
the site.
Site visit During the preliminary site visit, the ecological checklist was completed.
Most of the habitat is upland forest, old field, and successional terrestrial areas. Biological
surveys at this site have noted a variety of small mammal.signs. In addition, red-tailed hawks
were observed.
Problem formulation. At least 10 acres surrounding the site are known to be
contaminated with PCBs. Some PCBs are reproductive toxins in mammals (Ringer et al.,
1972; Aulerich et al., 1985; Wren, 1991; Kamrin and Ringer, 1996). When ingested, they
induce (i.e., increase concentrations and activity of) enzymes in the liver, which might affect
the metabolism of some steroid hormones (Rice and O'Keefe, 1995). Whatever the
mechanism of action, several physiological functions that are controlled by steroid hormones
can be altered by exposure of mammals to PCBs, and reproduction appears to be the most
sensitive endpoint for PCB toxicity in mammals (Rice and O'Keefe, 1995). Given this
information, the screening ecological risk assessment should include potential exposure
pathways for mammals to PCBs.,
Several possible exposure pathways were evaluated for mammals. PCBs are not
highly volatile, so inhalation of PCBs by animals would not be an important exposure
pathway. PCBs in soils generally are hot taken up by most plants, but are accumulated by
soil macroinvertebrates. Thus, herbivores, such as voles and rabbits, would not be exposed to
PCBs in most of their diets; whereas insectivores, such as shrews, or omnivores, such as deer
mice, could be exposed to accumulated PCBs in their diets. PCBs also are known to
biomagnify in terrestrial food chains; therefore, the ingestion exposure route needs evaluation,
and shrews and/or deer mice would be appropriate mammalian receptors to evaluate in this
exposure pathway.
Potential reproductive effects on predators that feed on shrews or mice also would be
important to evaluate. The literature indicated that exposure to PCBs through the food chain
could cause reproductive impairment in predatory birds through a similar mechanism as in
mammals. The prey of red-tail hawks include voles, deer mice, and various insects. Thus,
this raptor could be at risk of adverse reproductive effects.
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Ecological effects evaluation. No-observed-adverse-effect levels (NOAELs) for
the effects of PCBs and other contaminants at the site on mammals, birds, and other biota
were identified in the literature.
STEP 2: SCREENING-LEVEL EXPOSURE ESTIMATE AND RISK CALCULATION
Exposure estimate. For the screening-level risk calculation, the highest PCB and
other contaminant levels measured on site were used to estimate exposures.
Risk calculation. The potential contaminants of concern were screened based on
NOAELs for exposure routes appropriate to each contaminant. Based on this screen, PCBs
were confirmed to be the only contaminants of concern to small mammals, and possibly to
birds, based on the levels measured at this site. Thus, at the SMDP, the risk manager and
lead risk assessor decided to continue to Step 3 of the ecological risk assessment process.
STEP 3: BASELINE RISK ASSESSMENT PROBLEM FORMULATION
The screening-level ecological risk assessment confirmed that PCBs are of concern to
small mammals based on the levels measured at the site and suggested that predatory birds
might be at risk from PCBs that accumulate in some of their mammalian prey.
Ecotoxicity literature review. A literature review was conducted to evaluate
potential reproductive effects in birds. PCBs have been implicated as a cause of reduced.
reproductive success of piscivorous birds (e.g., cormorants, terns) in the Great Lakes (Kubiak
et al., 1989; Fox et al., 1991). Limited information was available on the effects of PCBs to
red-tailed hawks. A study on American kestrel indicated that consumption Of 33 mg/kgBW-
day PCBs resulted hi a significant decrease in sperm concentration in male kestrels (Bird et
al., 1983). Implications of this decrease for mating success in kestrels was not evaluated in
the study, but studies on other bird species indicate that it could increase the incidence of
infertile eggs and therefore reduce the number of young fledged per pair. The Great Lakes
International Joint Commission (IJC) recommends 0.1 mg/kg total PCBs as a prey tissue level
that will protect predatory birds and mammals (IJC, 1988). (This number is used as an
illustration and not to suggest that this particular level is appropriate for a given site.)
Exposure pathways. The complete exposure pathways identified during Steps 1
were considered appropriate for the baseline ecological risk assessment as well.
Assessment endpoints and conceptual model. Based on the screening-level
risk assessment for small mammals and the results of the ecotoxicity literature search for
birds, a conceptual model was initiated for the site, which included consideration of predatory
birds (e.g., red-tailed hawks) and their prey. The ecological risk assessor and the risk
manager agreed (SMDP) that assessment endpoints for the site would be the protection of
A-15
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small mammals and predatory birds from reproductive impairment caused by PCBs that had
accumulated in their prey.
An exposure pathway diagram was developed for the conceptual model to identify the
exposure pathways by which predatory birds could be exposed to PCBs originating in the soil
at the site (see Exhibit A-3). While voles may be prevalent at the site, they are not part of
the exposure pathway for predators because they are herbivorous and PCBs do not accumulate
in plants. Deer mice (Peromyscus maniculatus), on the other hand, also are abundant at the
site and, being omnivorous, are likely to be exposed to PCBs that have accumulated in the
insect component of their diet. Preliminary calculations indicated that environmental levels
likely to cause reproductive effects in predatory birds are lower than those likely to cause
reproductive effects in mice because mice feed lower in the food chain than do raptors. The
assessment endpoint was therefore restricted to reproductive impairment in predatory birds.
Risk questions. Based on the conceptual model, one question was whether
predatory birds could consume a high enough dose of PCBs in their diet to impair their
reproduction. Given the presence of red-tailed hawks on site, the question was refined to ask
whether that species could consume sufficient quantities of PCBs in their diet to affect
reproduction.
STEP 4: MEASUREMENT ENDPOINTS AND STUDY DESIGN
Measurement endpoints. To determine whether PCB levels in prey of the red-
tailed hawk exceed levels that might impair their reproduction, PCB levels would be
measured in deer mice taken from the site (of all of the species in the diet of the red-tailed
hawk, deer mice are assumed to accumulate the highest levels of PCBs). Based on estimated
prey ingestion rates for red-tailed hawks, a total PCB dose would be estimated from the
measured PCB concentrations in the mice.
Study design. The available measures of PCB concentrations in soil at the site
indicated a gradient of decreasing PCB concentration with increasing distance from the
unlined lagoon. Three locations along this gradient were selected to measure PCB
concentrations in deer mice. The study design specified that eight deer mice of the same size
and sex would be collected at each location. Each mouse should be approximately 20 grams
so that contaminant levels can be measured in individual mice. With concentrations measured
in eight individual mice, it is possible to estimate a mean concentration and an upper
confidence limit of the mean concentration in deer mice for the location. In addition, QA/QC
requirements dictate that an additional eight deer mice should be collected at one location.
For this site, it was necessary to verify that sufficient numbers of deer mice of the
specified size would be present to meet the sampling requirements. In addition, habitat
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EXHIBIT A-3
Conceptual Model for the Terrestrial PCB Site
MEASUREMENT ENDPOINT
(PCBs in mouse tissue, exposure
point for red-tailed hawks)
PRIMARY SOURCE
(Waste lagoons)
w
SECONDARY
SOURCE
(Site soils)
\
- h
w
PRIMARY RECEPTOR
(Deer mouse)
h
. ASSESSMENT
ENDPOINT-
SECONDARY RECEPTOR
(Predatory birds)
-------
conditions needed to be evaluated to determine what trapping techniques would work at the
targeted locations.
The study design specified further that the hazard quotient (HQ) method would be
used to estimate the risk of reproductive impairment in the red-tailed hawk from exposure to
PCBs in their prey. To determine the HQ, the measured DDT concentrations in deer mice is
divided by the LOAEL of 33 mg/kgBW-day for a decrease in sperm concentration in kestrels.
To estimate the dose to the red-tailed hawk, the PCB concentrations in deer mice is
multiplied by the quantity of deer mice that could be ingested by a red-tailed hawk each day
and divided by the body weight of the hawk. This dose is adjusted by a factor that
corresponds to the proportion of the diet of a red-tailed hawk that would come from the
contaminated area. This area use factor is a function of the home range size of the hawks
relative to the area of contamination. A HQ of greater than one implies that impacts due to
site contamination are likely, and an HQ of less than one implies impacts due to site
contaminants are unlikely.
STEP 5: FIELD VERIFICATION OF STUDY DESIGN
A field assessment using several trapping techniques was conducted to determine (1)
which technique was most effective for capturing deer mice at the site and (2) whether the
technique would yield sufficient numbers of mice over 20 grams to meet the specified
sampling design. On the first evening of the field assessment, two survey lines of 10 live
traps were set for deer mice in typical old-field habitat in the area believed to contain the
desired DDT concentration gradient for the study design. At the beginning of the second day,
the traps were retrieved. Two deer mice over 20 grams were captured in each of the survey
lines. These results indicated that collection of deer mice over a period of a week or less
with this number and spacing of live traps should be adequate,to meet the study objectives.
REFERENCES
Aulerich, R.J.; Bursian, SJ.; Breslin, WJ.; et al. 1985. Toxicological manifestations of
2,4,5-,2',4',5'-, 2,3,6,2',3',6'-, and 3,4,5,3',4',5'-hexachlorobiphenyl and Aroclor 1254
in mink. J. Toxicol. Environ. Health 15: 63-79.
Bird, D.M.; Tucker, P.H.; Fox, G.A.; Lague, P.C. 1983. Synergistic effects of Aroclor 1254
and mirex on the semen characteristics of American kestrels. Arch. Environ. Contain.
Toxicol. 12: 633-640.
Fox, G.A.; Collins, B.; Hayaskawa, E.; et al. 1991. Reproductive outcomes in colonial fish-
eating birds: a biomarker for developmental toxicants in Great Lakes food chains, n.
Spatial variation in the occurrence and prevalence of bill defects in young double-
crested cormorants in the Great Lakes. J. Great Lakes Res. 17:158-167.
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International Joint Commission (IJC) of United States and Canada. 1988. Great Lakes Water
Quality Agreement. Amended by protocol. Signed 18 November 1987. Ottawa,
Canada.
Kamrin, M.A.; Ringer, R.K. 1996. lexicological implications of PCB residues in mammals.
In: Beyer, W.N.; Heinz, G.H.; Redmon-Norwood, A.R. (eds.). Environmental
Contaminants in Wildlife: Interpreting Tissue Concentrations. A Special Publication
of the Society of Environmental Toxicology and Chemistry (SETAC), La Point, T.W.
(series ed.). Boca Raton, FL: CRC Press, Inc., Lewis Publishers, pp 153-164.
Kubiak, T.J.; Harris, H.J.; Smith, L.M.; et al. 1989. Microcontaminants and reproductive
impairment of the Forster's tem on Green Bay, Lake Michigan—1983. Arch. Environ.
Contam. Toxicol. 18: 706-727.
Rice, C.P; O'Keefe, P. 1995. Sources, pathways, and effects of PCBs, dioxins, and
dibenzofurans. In: Hoffman, D.J.; Rattner, B.A.; Burton, G.A. Jr.; Cairns, J., Jr.
(eds.). Handbook of Ecotoxicology. Ann Arbor, MI: CRC Press, Inc., Lewis
Publishers.
Ringer, R.K.; Aulerich, R.J.; Zabik, M. 1972. Effect of dietary polychlorinated biphenyls on
growth and reproduction of mink. Extended abstract. ACS (American Chemical
Society) 164th Annu. Meet. 12: 149-154.
Wren, C.D. 1991. Cause-effect linkages between chemicals and populations of mink
(M us tola vison) and otter (Lutra canadensis) in the Great Lakes basin. J. Toxicol.
Environ. Health 33: 549-585.
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APPENDIX B
REPRESENTATIVE SAMPLING GUIDANCE DOCUMENT,
VOLUMES: ECOLOGICAL
-------
OSWER Directive XXXXJDC
EPA 540/R/94/XXX
PBxx-xxxxxx
May 1997
DRAFT
SUPERFUND PROGRAM
REPRESENTATIVE SAMPLING GUIDANCE
VOLUMES: BIOLOGICAL
INTERIM FINAL
Environmental Response Team Center
Office of Emergency and Remedial Response
Office of Solid Waste and Emergency Response
U.S. Environmental Protection Agency
Washington. DC 20460
-------
Notice
The policies and procedures established in this document are intended solely for the guidance of government personnel, for
use in the Superfund Program. They are not intended, and cannot be relied upon, to create any rights, substantive or
procedural, enforceable by any party in litigation with the United States. The Agency reserves the right to act at variance
with these policies and procedures and to change them at any time without public notice.
For more information on Biological Sampling procedures, refer to the Compendium ofERT Toxicity Testing Procedures.
OSWER Directive 9360-44)8. EPA/540/P-91/009 (U.S. EPA 1991a). Topics covered in this compendium include: toxiciry
testing; and surface water and sediment sampling.
Please note mat the procedures in this document should only be used by individuals properly trained and certified under a
40 Hour Hazardous Waste Site Training Course that meets the requirements set forth in 29 CFR 1910.120(e)(3). It should
not be used to replace or supersede any information obtained in a 40 Hour Hazardous Waste Site Training Course.
Questions, comments, and recommendations are welcomed regarding the Superfund Program Representative Sampling
Guidance, Volume 3 -- Biological. Send remarks to:
Mark Sprenger PhD. - Environmental Scientist
David Charters Ph.D. - Environmental Scientist
U.S. EPA • Environmental Response Center (ERC)
Building 18, MS-101
2890 Woodbridge Avenue
Edison. NJ 08837-3679
For additional copies of the Superfund Program Representative Sampling Guidance, Volume 3 - Biological, contact*
National Technical Information Services
5285 Port Royal Road
Springfield. VA 22161
Phone (703) 487-4650
U.S. EPA employees can order a copy by calling the ERC at (908) 321-4212
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Disclaimer
This document has been reviewed in accordance with U.S. Environmental Protection Agency policy and approved for
publication. Mention of trade names or commercial products does not constitute endorsement or recommendation for use.
The following trade names are mentioned in this document:
Havahart® - Allcock Manufacturing Co.. Lititz, PA
Longworth - Longworth Scientific Instrument Company, Ltd.. England
Museum Special - Woodstream Corporation, Lititz, PA
Sherman - H.B. Sherman Traps, Tallahassee, FL
ui
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IV
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CONTENTS
Notice ; ii
Disclaimer iii
List of Figures viii
List of Tables : vii
Preface ;. ix
1.0 INTRODUCTION ; 1
1.1 Objective and Scope 1
1.2 Risk Assessment Overview 1
1.3 Conceptual Site Model 2
1.4 Data Quality Objectives 3
1.5 Technical Assistance 4
2.0 BIOLOGICAL/ECOLOGICAL ASSESSMENT APPROACHES 6
2.1 Introduction ...'... 6
2.2 RISK EVALUATION '. 6
2.2.1 Literature Screening Values 6
2.2.2 Risk Calculations 6
2.2.3 Standard Field Studies 6
2.2.3.1 Reference Area Selection 6
2.232 Receptor Selection 7
2.23.3 Exposure-Response Relationships • 8
2.23.4 Chemical Residue Studies 8
2.23.5 Population/Community Response Studies 9
2.2.3.6 Toxiciry Testing/Bioassays 9
3.0 BIOLOGICAL SAMPLING METHODS 11
3.1 Chemical Residue Studies 11
3.1.1 Collection Methods 11
3.1.1.1 Comparability Considerations 12
3.1.12 Mammals , i 12
3.1.1.3 Fish 13
3.1.1.4 Vegetation 13
3.1.2 Sample Handling and Preparation ' 14
3.1.3 Analytical Methods 14
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3.2 Population/community Response Studies 15
3.2.1 Terrestrial Vertebrate Surveys 15
3.2.2 Benthic Macroinvenebrate Surveys 15
3.2.2.1 Rapid Bioassessment Protocols for Benthic Communities 16
3.2.2.2 General Benthological Surveys 16
3.2.23 Reference Stations 16
3.2.2.4 Equipment for Benthic Surveys 16
3.2.3 Fish Biosurveys 17
3.23.1 Rapid Bioassessment Protocols for Fish Biosurveys 17
3.3 Tpxicity Tests 17
3.3.1 Examples of Acute Toxicity Tests •... 17
3.3.2 Examples of Chronic Toxicity Tests 18
4.0 QUALITY ASSURANCE/QUALITy CONTROL 21
4.1 . Introduction 21
4.2 Data Categories 21
4.3 Sources of Error 21
4.3.1 Sampling Design 21
4.3.2 Sampling Methodology and Sample Handling .. 22
4.3.3 Sample Homogeneity 22
4.3.4 Sample Analysis 22
4.4 QA/QC Samples i 23
4.4.1 Replicate Samples 23
4.4.2 Collocated Samples 24
4.43 Reference Samples 25
4.4.4 Rinsate Blank Samples 25
4.4.5 FieldBlank Samples .v 25
4.4.6 Trip Blank Samples 25
4.4.7 Performance Evaluation/Laboratory Control Samples 25
4.4.8 Controls ....'. 25
4.4.9 Matrix Spike/Matrix Spike Duplicate Samples 26
4.4.10 Laboratory Duplicate Samples , 26
4.5 Data Evaluation .26
4.5.1 Evaluation of Analytical Error 26
4.5.2 Data Validation 26
5.0 DATA ANALYSIS AND INTERPRETATION 27
5.1 Introduction 27
5.2 Data Presentation And Analysis 27
VI
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5.2.1 Data Presentation Techniques 27
522 Descriptive Statistics 27
5.23 Hypothesis Testing 27
5.3 Data Interpretation ; ;: 28
5.3.1 Chemical Residue Studies 28
5.3.2 Population/Community Studies .28
5.3.3 Toxicity Testing : 28
5.3.4 Risk Calculation 28
APPENDIX A - CHECKLIST FOR ECOLOGICAL ASSESSMENT/SAMPLING 30
APPENDK B - EXAMPLE OF FLOW DIAGRAM FOR CONCEPTUAL SITE MODEL 47
APPENDDC C - EXAMPLE SITES 50
REFERENCES 53
VII
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List of Figures
FIGURE 1 - Conceptual Site Model 5
FIGURE 2 - Common Mammal Traps 19
FIGURES - Illustrations of Sample Plots 29
List of Tables
TABLE 1 • Reference List of Standard Operating Procedures - Ecological Sampling Methods 20
vui
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Preface
This document is third in a series of guidance documents designed to assist Supcrfund Program Site Managers such as On-
Scene Coordinators (OSCs). Site Assessment Managers (SAMs). and other field staff in obtaining representative samples
at Supcrfund sites. It is intended to assist Supcrfund Program personnel in evaluating and documenting environmental threat
in support of management decisions, including whether or not to pursue a response action. This document provides general
guidance for collecting representative biological samples (i.e., measurement endpoints) once it has been determined by the
Site Manager that additional sampling will assist in evaluating the potential for ecological risk. In addition, this document
will:
• Assist field personnel in representative biological sampling within the objectives and scope of the Superfund
Program
• Facilitate the use of ecological assessments as an integral part of the overall site evaluation process
• Assist the Site Manager in determining whether an environmental threat exists and what methods are available to
assess that threat
This document is intended to be used in conjunction with other existing guidance documents, most notably. Ecological Risk -
Assessment Guidance for Superfund: Process for Designing and Conducting Ecological Risk Assessments, OSWER, EPA
540-R-97/006.
The objective of representative sampling is to ensure that a sample or a group of samples accurately characterizes site
conditions. Biological information collected in this manner complements existing ecological assessment methods.
Representative sampling within the objectives of the Superfund Program is used to:
promote awareness of biological and ecological issues
define the parameters of concern and the data quality objectives (DQOs)
develop a biological sampling plan
define biological sampling methods and equipment
identify and collect suitable quality assurance/quality control (QA/QQ samples
interpret and present the analytical and biological data
The National Contingency Plan (NCP) requires that short-term response (removal) actions contribute to the efficient
performance of any long-term site remediation, to the extent applicable. Use of this document will help determine if
biological sampling should be conducted at a site, and if so, what samples will assist program personnel in the collection
of information required to make such a determination.
j - '
Identification and assessment of potential environmental threats are important elements for the Site Manager to understand.
These activities can be accomplished through ecological assessments such as biological sampling. This document focuses
on the performance of ecological assessment screening approaches, more detailed ecological assessment approaches, and
biological sampling methods.
IX
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1.0 INTRODUCTION
1.1 OBJECTIVE AND SCOPE
This document is intended to assist Superfund Program
personnel in evaluating and documenting environmental
threat in support of management decisions. It presents
ecological assessment and sampling as tools in meeting
the objectives of trie Superfund Program, which include:
• ° Determine threat to public health, welfare, and
the environment
• Determine the need for long-term action
• Develop containment and control strategies
• Determine appropriate treatment and disposal
options
• Document attainment of clean-up goals
This document is intended to assist Superfund Program
personnel in obtaining scientifically valid and defensible
environmental data for the overall decision-making
process of site actions. Both the Comprehensive
Environmental Response. Compensation, and Liability
Act (CERCLA) [§104(a)(l)]. as amended by the
Superfund Amendments and Reauthorization Act
(SARA), and the NCP (§300.400(a)(2)]. require that the
United States Environmental Protection Agency (U.S.
EPA) "protect human health and the environment."
Environmental threats may. be independent of human
health threats, whether they co-exist at a site or are the
result of the same causative agents. It is therefore
important to determine and document potential.
substantial, and/or imminent threats to the environment
separately from threats to human health.
Representative sampling ensures that a sample or a group
of sample accurately characterizes site conditions.
Representative biological sampling and ecological risk
assessment include, but are not limited to. the collection
of sue information and the collection of samples for
chemical or lexicological analyses. Biological sampling
is dependent upon specific site requirements during
limited response actions or in emergency response
situations. Applying the methods of collecting
environmental information, as outlined in this document,
can facilitate the decision-making process (e.g., during
chemical spill incidents).
The collection of representative samples is critical to the
site evaluation process since all data interpretation
assumes proper sample collection. Samples collected
which inadvertently or intentionally direct the generated
data toward a conclusion are biased and therefore not
representative.
This document provides Superfund Program personnel
with general guidance for collecting representative
biological samples (i.e., measurement endpoints, [see
Section 1.2 for the definition of measurement endpoint]).
Representative biological sampling is conducted once the
Site Manager has determined that additional sampling
may assist in evaluating the potential for ecological risk.
This determination should be made in consultation with
a trained ecologist or biologist The topics covered in
this document include sampling methods and equipment,
QA/QC, and data analysis and interpretation.
The appendices in this document provide several types of
assistance. Appendix A provides a checklist for initial
ecological assessment and sampling. Appendix B
provides an example flow diagram for the development
of a conceptual site model. Appendix C provides
examples of how the checklist for ecological
assessment/sampling is used to formulate a conceptual
site model that leads up to the design of a site
investigation.
This document is intended to be used in conjunction with
other existing guidance documents, most notably,
Ecological Risk Assessment Guidance for Superfund:
Process for Designing and Conducting Ecological Risk
Assessments, EPA 540-R-97/006 (U.S. EPA 1997).
1.2 RISK ASSESSMENT OVERVIEW
The term ecological risk assessment (ERA), as used in
this document, and as defined in Ecological Risk
Assessment Guidance for Superfund: Process for
Designing and Conducting Ecological Risk
Assessments, OSWER, EPA 540-R-97/006 (U.S. EPA
1997) refers to:
"... a qualitative and/or quantitative
appraisal of the actual or potential
impacts of a hazardous waste site on
plants and animals other than humans
and domesticated species."
Risk assessments are an integral pan of the Superfund
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process and are conducted as part of the baseline risk
assessment for the remedial investigation and feasibility
study (RI/FS). The RI is defined by a characterization of
the nature and extent of contamination, and ecological
and human health risk assessments. The nature and
extent of contamination determines the chemicals present
on the site. The ecological and human health risk
assessments determine if the concentrations threaten the
environment and human health.
An ecological risk assessment is a formal process that
integrates knowledge about an environmental
contaminant (i.e., exposure assessment) and its potential
effects to ecological receptors (i.e., hazard assessment).
The process evaluates the likelihood that adverse
ecological effects may occur or are occurring as a result
of exposure to a stressor. As defined by U.S. EPA
(1992), a stressor is any physical, chemical or biological
entity that can induce an adverse ecological response.
Adverse responses can range from sublethal chronic
effects in an individual organism to a loss of ecosystem
function.
Although stressors can be biological (e.g., introduced
species), in the Superfund Program substances
designated as hazardous under CERCLA are usually the
stressors of concern. A risk does not exist unless (1) the
stressor has the ability to cause one or more adverse
effects, and (2) it co-occurs with or contacts an ecological
component long enough and at sufficient intensity to elicit
the identified adverse effect.
The risk assessment process also involves the
identification of assessment and measurement endpoints.
Assessment endpoints are explicit expressions of the
actual environmental values (e.g., ecological resources)
that are to be protected. A measurement endpoint is a
measurable biological response to a stressor that can be
related to the valued characteristic chosen as the
assessment endpoint (U.S. EPA 1997). Biological
samples are collected from a site to represent these
measurement endpoints. See Section 2.2 for a detailed
discussion of assessment and measurement endpoints.
Except where required under other regulations, issues
such as restoration, mitigation, and replacement are;
i important to the program but are reserved for:
; investigations that may or may not be included in the RI i
I phase. During the management decision process of:
i selecting the preferred remedial option leading to the
! Record of Decision (ROD), mitigation and restoration
i issues should be addressed. Note that these issues are not j
i necessarily issues within the baseline ecological risk I
assessment.
Guidelines for human health risk assessment have been
established; however, comparable protocols for
ecological risk assessment do not currently exist.
Ecological Risk Assessment Guidance for Superfund:
Process for Designing and Conducting Ecological Risk
Assessments." (U.S. EPA 1997) provides conceptual
guidance and explains how to design and conduct
ecological risk assessments for a CERCLA RI/FS. The
Framework for Ecological Risk Assessment (U.S. EPA
1992) provides an Agency-wide structure for conducting
ecological risk assessments and describes the basic
elements for evaluating site-specific adverse effects of
stressors on the environment- These documents should
be referred to for specific information regarding the risk
assessment process.
While the ecological risk assessment is a necessary first
step in a "natural resource damage assessment" to
provide a causal link, it is not a damage evaluation. A
natural resource damage assessment may be conducted at
any Superfund site at the discretion of the Natural
Resource Trustees. The portion of the damage
assessment beyond the risk assessment is the
responsibility of the Natural Resource Trustees, not of the
U.S. EPA. Therefore, natural resource damage
assessment is not addressed in this guidance.
1.3 CONCEPTUAL SITE MODEL
A conceptual site model is an integral pan of a site
investigation and/or ecological risk assessment as it
provides the framework from which the study design is
structured. The conceptual site model follows
contaminants from their sources, through transport and
fate pathways (air, soil, surface water, groundwater), to
the ecological receptors. The conceptual model is a
strong tool in the development of a representative
sampling plan and is a requirement when conducting an
ecological risk assessment. It assists the Site Manager in
evaluating the interaction of different site features (e.g.,
drainage systems and the surrounding topography),
thereby ensuring that contaminant sources, pathways, and
ecological or human receptors throughout the site have
been considered before sampling locations, techniques,
and media are chosen.
Frequently, a conceptual model is created as a site map
(Figure 1) or flow diagram that describes the potential
movement of contaminants to site receptors (see
Appendix B). Important considerations when creating a
conceptual model are:
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• The state(s) (or chemical form) of each
contaminant and its potential mobility through
various media
• Site topographical features
• Meteorological conditions (e.g., climate,
precipitation, humidity, wind direction/speed)
• Wildlife area utilization.
Preliminary and historical site information may provide
the identification of the contaminants) of concern and the
tevel(s) of the contamination. A sampling plan should be
developed from the conceptual model based on the
selected assessment endpoints.
The conceptual site model (Figure 1) is applied to this
document, Representative Sampling Guidance Volume
3: Biological. Based on the model, you can
approximate:
• Potential Sources
hazardous waste site (waste pile, lagoon,
emissions), drum dump (runoff, leachate),
agricultural (runoff, dust, and particulates)
• Potential Exposure Pathways
ingestion
waste, contained -in the pile on the
hazardous waste site; soil panicles near
the waste pile; drum dump; or area of
agricultural activity
inhalation
dust and particulates from waste pile,
drum dump, or area of agricultural activity
absorption/direct contact
soil near waste pile, drum dump, or area of
agricultural activity and surface water
downstream of sources
• Potential Migration Pathways
air (paniculates and gases) from drum dump
and area of agricultural activity
soil (runoff) from the hazardous waste site.
drum dump, and agricultural runoff
surface water (river & lake) from hazardous
waste site and agricultural runoff
groundwater (aquifer) from drum dump
leachate.
• Potential Receptors of Concern (and associated
potential routes)
wetland vegetation/mammals/invertebrates if
suspected to be in contact with potentially
contaminated soil and surface water
rivenne vegetation/aquatic organisms if
suspected to be in contact with potentially
contaminated surface water and soil
lake vegetatum/mammals/aquatic organisms if
suspected to be in contact with potentially
contaminated surface water and leachate.
1.4 DATA QUALITY OBJECTIVES
Data quality objectives (DQOs) state the level of
uncertainty that is acceptable from data collection
activities. DQOs also define the data quality necessary to
make a certain decision. Consider the following when
establishing DQOs for a particular project:
• Decision(s) to be made or question(s) to be
answered;
• Why environmental data are needed and how
the results will be used;
• Time and resource constraints on data
collection;
• Descriptions of the environmental data to be
collected;
• Applicable model or data interpretation method
used to arrive at a conclusion;
• Detection limits for analytes of concern; and
• Sampling and analytical error.
In addition to these considerations, the quality assurance
components of precision, accuracy (bias), completeness.
representativeness, and comparability should also be
considered. Quality assurance components are defined as
follows:
• Precision - measurement of variability in the
data collection process.
• Accuracy (bias) - measurement of bias in the
analytical process. The term "bias" throughout
this document refers to the QA/QC accuracy
component.
• Completeness — percentage of sampling
measurements which are judged to be valid.
• Representativeness - degree to which sample
data accurately and precisely represent the
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characteristics of the site contaminants and their
• . concentrations.
• Comparability — evaluation of the similarity of
conditions (e.g., sample depth, sample
homogeneity) under which separate sets of data
are produced.
Many of the DQOs and quality assurance considerations
for soil, sediment, and water sampling are also applicable
to biological sampling. However, there are also
additional considerations that are specific to biological
sampling.
• Is biological data needed to answer the
question(s) and, if so, how will the data be used;
• Seasonal, logistical, resource, and legal
constraints on biological specimen collection;
• What component of the biological system will
be collected or evaluated (i.e., tissue samples.
whole organisms, population data, community
data, habitat data);
• The specific model or interpretation scheme to
be utilized on the data set;
• The temporal, spatial, and behavioral variability
inherent in natural systems.
Quality assurance/quality control (QA/QC) objectives are
discussed further in Chapter 4.
1.5 TECHNICAL ASSISTANCE
In this document, ii is assumed that technical specialists
are available to assist Site Managers and other site
personnel hi determining the best approach to ecological
assessment. This assistance ensures that all approaches
are up-to-date and that best professional judgment is
exercised. Refer to Appendix A for more information.
Support in designing and evaluating ecological
assessments is currently available from regional technical
assistance groups such as Biological Technical
Assistance Groups (BTAGs). Support is also available
from the Environmental Response Team Center (ERTC)
as well as from other sources within each region.
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FIGURE 1:
CONCEPTUAL
SITE MODEL
WIND ROSE
STATE GAME
LANDS
syN. r N5j*iL3|S*&
^\%&l
WATER PLANT
INTAKE
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2.0 BIOLOGICAL/ECOLOGICAL ASSESSMENT APPROACHES
2.1 INTRODUCTION
Biological assessments vary in their level of effort,
components, and complexity, depending upon the
objectives of the study and specific site conditions. An
assessment may consist of literature-based risk
evaluations and/or site-specific studies (e.g.,
population/community studies, toxicity tests/bioassays,
and tissue residue analyses).
Superfund Program personnel (RPMs and OSCs) may be
limited to completing the ecological checklist (Appendix
A) during the. Preliminary Site Evaluations and to
consulting an ecological specialist if it is determined that
additional field data are required. The checklist is
designed to be completed by one person during an initial
-.site visit. The checklist provides baseline data, is useful
in designing sampling objectives; and requires a few
hours to complete in the field.
When the Site Manager determines that additional data
collection is needed at a response site, the personnel and
other resources required depends on the selected
approach and the site complexity.
To determine which biological assessment approach or
combination of approaches is appropriate for a given site
or situation, several factors must be considered. These
include what management decisions will ultimately need
to be made based on the data; what are the study
objectives: and what should be the appropriate level of
effort to obtain knowledge of contaminant fate/ transport
and ecotoxicity.
2.2 RISK EVALUATION
Three common approaches to evaluating environmental
risk to ecological receptors are (1) the use of literature
screening values (e.g.. literature toxicity values) for
comparison to site-specific contaminant levels. (2) a
"desk-top" risk assessment which can model existing site-
specific contaminant data to ecological receptors for
subsequent comparison to literature toxicity values, and
(3) field investigation/laboratory analysis that involves a
site investigation (which may utilize existing contaminant
data for support) and laboratory analysis of contaminant
levels in media and/or experimentation using bioassay
procedures. These three approaches are described in
further detail next.
2.2.1 Literature Screening Values
To determine the environmental effects of contaminants
at a hazardous waste site, the levels of contaminants
found may be compared to literature toxicity screening
values or established screening criteria. These values
should be derived from studies that involve testing of the
same matrix and a similar organism of concern. Most
simply stated, if the contaminant levels on the site are
above the established criteria, further evaluation of the
site may be necessary to determine the presence of risk.
Site contaminant levels that are lower than established
criteria may indicate that no further evaluation is
necessary at the site for that contaminant
2.2.2 Risk Calculations
The "desk-top" risk calculation approach compares site
contaminants to information from studies found in
technical literature. This type of evaluation can serve as
a screening assessment or as a tier in a more complex
evaluation. Since many assumptions must be made due
to limited site-specific information, risk calculations are
necessarily conservative. The collection and inclusion of
site-specific field data can reduce the number and/or the
magnitude of these "conservative" assumptions, thereby
generating a more realistic calculation of potential risk.
(See Chapter 5.0 for a complete discussion on risk
calculations.)
2.2.3 Standard Field Studies
Two important aspects of conducting a field study that
warrant discussion are the selection of a reference area
and the selection of the receptors of concern. These are
important to establish prior to conducting a field study.
2.2.3.1 Reference Area Selection
A reference area is defined in this document as an area
that is outside the chemical influence of the site but
possesses similar characteristics (e.g., habitat, substrate
type) that allows for the comparison of data between the
impacted area (Le., the site) and the unimpacted area (i.e.,
the reference area). Reference areas can provide
information regarding naturally occurring compounds and
the existence of any regional contamination independent
of the site. They can help determine if contaminants are
ubiquitous in the area and can separate site-related issues
-------
from non-she related issues.
The reference area must be of similar habitat type and
support a species composition similar to the study area.
The collection and analysis of samples from a reference
area can support site-specific decisions regarding uptake,
body burden, and accumulation of chemicals and toxicity.
The reference area should be outside the area of influence
of the site and if possible, in an area of minimal
contamination or disturbance. Location of reference
areas in urban or industrial areas is frequently difficult,
but an acceptable reference area is usually critical to the
successful use of ecological assessment methods.
2.2.3.2 Receptor Selection
The selection of a receptor is dependent upon the
objectives of the study and the contaminants present The
. first step is to determine the toxicity characteristics of the
contaminants (i.e., acute, chronic, bioaccumulative, or
non-persistent). The next step is to determine the
exposure route of the chemical (i.e., dermal, ingestion.
inhalation).
Selection of the receptor or group of receptors is a
component of establishing the measurement endpoint in
the study design. When discussing the term measurement
endpoint, it is useful to first define a related concept, the
assessment endpoint. An assessment endpoint is defined
as "an explicit expression of the environmental value that
is to be protected." For example, "maintaining aquatic
community composition and structure downstream of a
site similar to that upstream of the site" is an explicit
assessment endpoint. Inherent in this assessment endpoint
is the process of receptor selection that would most
appropriately answer the question that the endpoint
raises. Related to this assessment endpoint is the
measurement endpoint which is defined as "a measurable
ecological characteristic that is related to the valued
characteristic chosen as the assessment endpoint." For
'example, measurements of biological effects such as
mortality, reproduction, or growth of an invertebrate
community are measurement endpoints. Establishing
these endpoints will ensure (1) that the proper receptor
will be selected to best answer the questions raised by the
assessment and measurement endpoints, and (2) that the
focus of the study remains on the component of the
environment that may be used as the basis for decision.
There are a number of factors that must be considered
when selecting a target species. The behavioral habits
and lifestyle of the species must be consistent with the
environmental fate and transport of the contaminants of
interest as well as pathways of exposure to receptor
species. For example, if the contaminants of concern at
the site are PCBs that are bioaccumulative, a mammal
such as a mink could be selected for the study since this
species is documented to be sensitive to the
bioaccumulation of PCBs. The r"'n^ in this case has
been selected to be used far establishing the measurement
endpoint that is representative of piscivorous mammals.
However, h may not be feasible to collect mink for study
due to their low availability in a given area. Therefore.
the food items of the mink (e.g., small tnammnic aquatic
vertebrates and invertebrates) may be collected and
analyzed for PCBs as an alternative means of evaluating
the risk to mink. The resulting residue data may be
utilized to produce a dose model. From this model, a
reference dose value may be determined from which the
probable effects to mink calculated.
The movement patterns of a measurement endpoint are
also important during the receptor selection process.
Species that are migratory or that have large feeding
ranges are more difficult to link to site exposure than
those which are sessile, territorial, or have limited
movement patterns.
Ecological field studies offer direct or corroborative
evidence of a link between contamination and ecological
effects. Such evidence includes:
• Reduction in population sizes of species that
can not be otherwise explained by naturally
occurring population cycles
• Absence of species-normally occurring in the
habitat and geographical distribution
• Dominance of species associated primarily with
stressed habitat
• Changes in community diversity or trophic
structure relative to a reference location
• High incidence of lesions, tumors, or other
pathologies
• Development of exposure response
relationships.
Ecologists usually compare data of observed adverse
effects to information obtained from a reference area not
affected by site contamination. To accomplish this.
chemical and biological data should be collected
simultaneously and then compared to determine if a
correlation exists between contaminant concentrations
and ecological effects (U.S. EPA 1991b). The
simultaneous collection of the data is important in
reducing the effect of temporal variability as a factor in
the correlation analysis.
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The type of field study selected is directed by the
contaminants present linked to the assessment endpoint.
Prior to choosing a specific study approach, the site
contaminant must be determined using information about
known or suspected site contaminants and how the nature
of these contaminants may be modified by several
environmental and ecotoxicological factors. In addition,
evaluation of chemical fate and transport information is
necessary to determine the appropriate matrix and
technique.
Contaminants can be a food chain threat, a lethal threat,
a direct non-lethal toxicant, indirect toxicant, or some
combination of the four. Chemical residue studies are
appropriate if the contaminant of concern (COC) will
bioaccumulate. Ecotoxicological information can provide
insight about contaminants that are expected to
accumulate in organisms. It can also provide information
about which organisms provide the best data for the study
objectives. For example, the species-specific
bioaccumulation rate must be considered along with
analytical detection limits; the bioaccumulated levels
need to be above the analytical detection limits. In
contrast, population/ community studies or toxicity testing
may be more appropriate if the contaminants cause direct
lethality.
2.2.3.3 Exposure - Response Relationships
The relationship between the exposure (or dose) of a
contaminant and the response that it elicits is a
fundamental concept in toxicology (Timbrell 1989). The
simplest response to observe is death. Some examples of
other responses that vary in terms of ease of measurement
include pathological lesions, cell necrosis, biochemical
changes, and behavioral changes. It is this foundation of
exposure-response relationships upon which the concept
of chemical residue studies, population/community
studies, and toxicity lesting/bioassays are built upon.
2.2.3.4 Chemical Residue Studies
Residue studies are appropriate to use when there is
concern about the accumulation of contaminants in the
tissues of indigenous species. Residue studies are
conducted by collecting organisms of one or more species
and comparing the contaminant bioaccumulation data to
those organisms collected from a reference area.
Chemical residue studies require field collection of biota
and subsequent tissue analysis. A representative
organism for collection and analysis is selected based on
the study objectives and the site habitat. Generally the
organism should be abundant, sessile (or with limited
home range), and easy to capture. These attributes help
to provide a sufficient number of samples for analysis
thereby strengthening the linkage to the site. A number
of organism- and contaminant-specific factors should also
be considered when designing residue studies (see Philips
[1977] and [1978] for additional information). The
subsequent chemical analysis may be conducted on
specific target tissues or the whole body. In most cases,
whole-body analysis is the method of choice to support
biological assessments. This is because most prey
species are eaten in entirety by the predator.
In designing residue analysis studies, it is important to
evaluate the exposure pathway carefully. If the organisms
analyzed are not within the site-specific exposure
pathway, the information generated will not relate to the
environmental threat Evaluation of the exposure
pathway may suggest that a species other than the one of
direct concern might provide a better evaluation of
potential threat or bioaccumulation.
Because there are different data needs for each objective,
the study objective needs to be determined prior to the
collection of organisms. In these studies the actual
accumulation (dependent upon the bioavailability) of the
contaminants is evaluated rather than assumed from
literature values. The information collected then allows
for site-specific evaluation of the threat and reduces the
uncertainty associated with the use of literature
bioavailability values. These factors may be applied for
specific areas of uncertainty inherent from the
extrapolation of available data (e.g., assumptions of 100
percent bioaccumulation, variations in sensitive
populations).
As stated previously, because site conditions as well as
the bioavailability can change over time, it is important
that exposure medium (soil, sediment, or water) samples
and biological samples are collected simultaneously and
analyzed for the same parameters to allow for the
comparison of environmental contaminant levels in the
tissue and the exposure medium. This is critical in
establishing a site-specific linkage that must be
determined on a case-by-case basis.
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2.2.3.5 Population/Community
Response Studies
The fundamental approach to population or community
response studies is to systematically sample an area,
documenting the organisms of the population or
community. Individuals are typically identified and
enumerated, and calculations are made with respect to the
number, and species present. These calculated values
(e.g., indices or metrics) are used to compare sampling
locations and reference conditions. Some population and
community metrics include the number of individuals,
species composition, density, diversity, and community
structure.
2.2.3.6 Toxicity Testing/Bioassays
A third common assessment approach is to utilize toxicity
tests or bioassays. A toxicity test may be designed to
measure the effects from acute (short-term) or chronic
(long-term) exposure to a contaminant. An acute test
attempts to expose the organism to a stimulus that is
severe enough to produce a response rapidly. The
duration of an acute toxicity test is short relative to the
organism's life cycle and mortality is the most common
response measured. In contrast, a chronic test attempts
to induce a -biological response of relatively slow
progress through continuous, long-term exposure to a
contaminant.
In designing a toxicity test, it is critical to understand the
fate, transport, and mechanisms of toxicity of the
contaminants to select the test type and conditions. The
toxicity test must be selected to match the site and its
conditions rather than modify the site matrix for the use
of a particular test. Factors to consider are the test
species, physical/chemical factors of the contaminated
media, acclimation of test organisms, necessity for
laboratory versus field testing, test duration, and selection
of test endpomis (e.g.. mortality or growth). A thorough
understanding of the interaction of these and other factors
is necessary to determine if a toxicity test meets the study
objectives.
The selection of the best toxicity test, including the choice
of test organism, depends on several factors:
• The decisions that will be based on the results
of the study
• The ecological setting of the site
• The contaminant(s) of concern
Toxicity testing can be conducted on a variety of sample
matrices, including water (or an aqueous effluent),
sediment, and soil. Soil and sediment toxicity tests can
be conducted on the parent material (solid-phase tests) or
on the elutriate (a water extract of the soil or sediment).
Solid-phase sediment and soil tests are currently the
preferred tests since they evaluate the toxicity of the
matrix of interest to the test organisms, thereby providing
more of a realistic site-specific exposure scenario.
As stated previously, one of the most frequently used
endpoints in acute toxicity testing is mortality (also
refaied to as lethality) hmnter ft is one of the most easily
measured parameters.
In contrast, some contaminants do not cause mortality in
; but rather they affect the rate or success of
reproduction or growth in test organisms. In this case,
the environmental effect of a contaminant may be that it
causes reproductive failure but does not cause mortality
in the existing population. In either case, the population.
will either be eliminated or drastically reduced.
The use of control as well as reference groups is normally
required. Laboratory toxicity tests include a control that
evaluates die laboratory conditions, and the health and
response of the test organisms. Laboratory controls are
required for all valid toxicity tests. A reference provides
information on how the test organisms respond to the
exposure medium without the site contaminants.
Therefore, the reference is necessary for interpretation of
the test results in the context of the site (i.e., sample data
is compared to the reference data). It is not uncommon
for conditions other than contamination to induce a
response in a toxicity test With proper reference and
control tests, toxicity tests can be used to establish a link
between contaminants results and adverse effects.
Within the Superfund Program, conducting toxicity tests
typically involves collecting field samples (water.
sediment, soil) and transferring the materials to a
laboratory. In situ (field conducted) tests can be run if
field conditions permit There are benefits and
limitations associated with each approach. The most
notable benefit of laboratory testing is that exposure
conditions are controlled, but mis leads to its most
notable limitation, a reduction of realism. With in situ
tests, the reality of the exposure situation is increased, but
there is a reduction of test controls. See U.S. EPA's
Compendium of ERT Toxicity Testing Procedures,
OSWER Directive 9360.44)8, EP A/54 O/P-91/009 (U.S.
EPA 199la), for descriptions of nine common toxicity
tests and Standard Guide for Conducting Sediment
Toxicity Tests with Freshwater Invertebrates, ASTM
Standard E1383. October 1990.
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Species Seleetipn for Toxieitv Testing
Selection of the test organism is critical in designing a
study using toxicity testing. The species selected should
be representative relative to the assessment endpoint,
typically an organism found within the exposure pathway
expected in the field. To be useful in evaluating risk, the
test organism must respond to the contaminant(s) of
concern. This can be difficult to achieve since the species
and tests available are limited. Difficult choices and
balancing of factors are frequently necessary.
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3.0 BIOLOGICAL SAMPLING METHODS
Once a decision has been made that additional data are
required to assess the biological threat posed by a site, an
appropriate sampling plan must be developed. The
selection of ecological sampling methods and equipment
is dependent upon the field assessment approach, as
discussed in Chapters 1 and 2. Thus, the selection of an
assessment approach is the initial step in the collection
process. This chapter does not present step-by-step
instructions for a particular method, nor does it present an
exhaustive list of methods or equipment Rather, it
presents specific examples of the most commonly used
methods and associated equipment. Table 4.1 (at the end
of this chapter) lists some of the standard operating
procedures (SOPs) used by the U.S. EPA's
Environmental Response Team Center (ERTC).
.Because of the complex process required for selecting
the proper assessment approach for a particular site,
consultation with an ecologist/biologist experienced in
conducting ecological risk assessments is strongly
recommended.
3.1 CHEMICAL RESIDUE STUDIES
Chemical residue studies are a commonly used approach
that can address the bioavailability of contaminants in
media (e.g., soil, sediment, water). They are often called
tissue residue studies because they measure the
contaminant body burden in site organisms.
When collecting organisms for tissue analyses, it is
critical that the measured levels of contaminants in the
organism are attributable to a particular location and
contaminant level within the site. Collection techniques
must be evaluated for their potential to bias the generated
data. Collection methods can result in some form of
biased data either by the size. sex. or individual health of
the organism. Collection techniques are chosen based on
the habitat present and the species of interest. When
representative approaches are not practical, the potential
bias must be identified and considered when drawing
conclusions from the data. The use of a particular
collection technique should not be confused with the need
to target a "class" of individuals within a population for
collection. For example, in a specific study it may be
desirable to collect only males of the species or to collect
fish of consumable size.
Some receptors of concern (ROCs) cannot be collected
and analyzed directly because of low numbers of
individuals in the study area, or other technical or
logistical reasons. Exposure levels for these receptors
can be estimated by collecting organisms that are preyed
upon by the ROC. For example, if the ROC is a
predatory bird, the species collected for contaminant level
measurements may be one of several small mammals or
fish that the ROC is known to eat
As noted previously, it is critical to link the accumulated
contaminants both to the site and to an exposure medium.
Subsequently, the collection and analysis of
representative soil, sediment, or water samples from the
same location are critical. A realistic site-specific
Bioaccumulation Factor (BAF) or Biocohcentration
Factor (BCF) may then be calculated for use in the site
exposure models.
"Bioconcentration is usually considered to be that process
by which toxic substances enter aquatic organisms, by
gill or epithelial tissue from the water. Bioaccumulation
is a broader term in the sense that it usually includes not
only bioconcentration but also any uptake of toxic
substances through the consumption of one organism."
(Brungs and Mount 1978).
3.1.1 Collection Methods
It should be noted that any applicable state permits
should be acquired before any biological sampling event.
States requirements on organism, method, sampling
location, and data usage differ widely and may change
from year to year.
The techniques used to collect different organisms are
specific to the study objectives. All techniques are
selective to some extent for certain species, sizes, habitat,
or sexes of animals. Therefore, the potential biases
associated with each technique should be determined
prior to the study. If the biases are recognized prior to
collection, the sampling may be designed to minimize
effect of the bias. For example, large traps are not
effective for trapping small animals since small mammals
are not heavy enough to trigger the trap or may escape
through minute trap openings.
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In determining environmental threat, the target species
generally consist of prey species such as earthworms,
small mammals, or fish. Residue data from these
organisms can be used to evaluate the risk to higher
trophic level organisms, which may be difficult to capture
or analyze.
3.1.1.1 Comparability Considerations
There are two issues that directly affect field collection.
First, organisms such as benthic macroinvenebrates tend
to have a patchy or non-uniform distribution in the
environment due to micro habitats and other factors.
Therefore, professional evaluation in matching habitat for
sampling, is critical in the collection of * a truly
representative sample of the community. Second,
variability in sampling effort and effectiveness needs to
be considered.
3.1.1.2 Mammals
Trapping is the most common method for the collection
of mammals. The selection of traps is determined by the
species targeted and the habitat present Both live trap or
kill trap methods may be acceptable for residue studies,
but consideration of other data uses (e.g., histopathology)
or concern for injury or death of non-target species can
influence the use of certain trap types.
Several trap methods are available for collecting small
mammals. Commonly used traps include Museum
Special. Havahart Longworth. and Sherman traps
(Figure 3). Although somewhat labor-intensive, pitfall
trap arrays may also be established to include mammals
that are not regularly trapped using other techniques (e.g.,
shrews).
Trap placement is a key element when collecting
samples. Various methods of trap placement can be
utilized. These include, but are not limited to:
• Sign method/Best set method
• Paceline method
• Grid method
When using the sign/best set method, an experienced
field technical specialist searches for fresh mammal signs
(e.g.. tacks, scat, feeding debris) to determine'where the
trap should be positioned. This method typically
produces higher trapping success than other methods,
however, this method is biased and is therefore generally
used to determine what species are present at the site.
The paceline method involves placement of traps at
regular intervals along a transect. A starting point is
selected and marked, a landmark is identified to indicate
the direction of the transect, and as the field member
walks the. transect, the traps are placed at regular
intervals along it
The grid method is similar to the paceline method but
involves a group of evenly spaced parallel transects of
equal lengths to create a grid. Traps are placed at each
grid node. The size of the grid is dependent on the
species to be captured and the type of study. Grids of
between 500 to 1,000 square meters containing
approximately 100 traps are common. If a grid is
established in a forest interior, additional parallel
trapping lines may be established to cover the edge
habitat
Regardless of the type of .trapping used, habitat
disturbance should be kept to a minimum to achieve
maximum trapping success. In most areas, a trapping
success of 10 percent is considered maximum but is
oftentimes significantly lower (e.g., 2 to 5 percent). Part
of this reduced trapping success is due to habitat
disturbance. Therefore, abiotic media samples (e.g., soil,
sediment, water) should be collected well in advance of
trapping efforts or after all trapping is completed.
Trapping success also varies with time but may increase
over time with diminishing returns. In other words.
extending the trapping period over several days may
produce higher trapping success by allowing mammals
that were once peripheral to the trapping area to
immigrate into the now mammal-depauperate area.
These immigrants would not be representative of the
trapping area. Therefore, a trapping period of 3 days is
typically used to minimize this situation.
Trapping success will also vary widely based on the
available habitat, targeted species, season, and
geographical location of the site. When determining trap
success objectives, it is important to keep in mind the
minimum sample mass/volume requirements for chemical
residue studies.
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•3.1.1.3. Fish
Electrofishing, gill nets, trawl nets, seine nets, and
minnow traps are common methods used for the
collection of fish. The selection of which technique to
use is dependent on the species targeted for collection
and the systerh being sampled. In addition, there are
other available fish netting and trapping techniques that
may be more appropriate in specific areas. As with
mammal trapping, disturbance in the area being sampled
should be kept to a minimum to ensure collection
success.
Electrofishing uses electrical currents to gather, slow
down, or immobilize fish for capture. An electrical field
is created between and around two submerged electrodes
that stuns the fish or alters their swimming within or
around the field Depending on the electrical voltage, the
electrical pulse frequency, and the fish species, the fish
may swim towards one of the electrodes, swim slowly
enough to capture, or may be stunned to the point of
immobilization. This technique is most effective on fish
with swimbladders and/or shallow water since these fish
will float to the surface for easy capture.
Electrofishing can be done using a backpack-mounted
electroshocker unit, a shore-based unit or from a boat
using either type. Electrofishing does not work in saline
waters and can be ineffective in very soft water.
Electrofishing is less effective in deep water where the
fish can avoid the current. In turbid waters, it may be
difficult to see the stunned fish.
Gill netting is a highly effective passive collection
technique for a wide range of habitats. Because of its low
visibility under water, a gill net captures fish by
entangling their gill plates as they attempt to swim
through the area in which the gill net has been placed in.
Unfortunately, this may result in fish to be injured or
killed due to further entanglement, predation, or fatigue.
The size and shape of fish captured is relative to the size
and kind of mesh used in the net thus creating bias
towards a certain sized fish. These nets are typically used
in shallow waters, but may extend to depths exceeding SO
meters. The sampling area should be free of obstructions
and floating debris, and provide little to no current.
(Hurben 1983)
Otter trawl netting is an active collection technique that
utilizes the motion of a powered boat to drag a pocket-
shaped net through a body of water. The net is secured to
the rear of a boat and pulled to gather any organisms that
are within the opening of the pocket This pocket is kept
open through the use of underwater plates on either side
of the net that act as keels, spreading the mouth of the net
open.
Seining is another active netting technique that traps fish
by encircling mem with a long wall of netting. The top of
the net is buoyed by floats and the bottom of the net is
weighed down by lead weights or chains. Seine nets are
effective in open or shallow waters with unobstructed
bottoms. Beach or haul seines are used in shallow water
situations where the net extends to the bottom. Purse
seines are designed for applications in open water and do
not touch the bottom (Hayes 1983).
The use of minnow traps is a passive collection technique .
for minnow-sized fish. The trap itself is a metal or plastic
cage that is secured to a stationary point and baited to
attract fish. Small funnel-shaped openings on either end
of the trap allow fish to swim easily into it, but are
difficult to locate for exit Cage "extenders" or "spacers"
that are inserted to lengthen the cage, allow larger
organisms such as eels, or for a larger mass offish to be
collected.
3.1.1.4 Vegetation
Under certain conditions, the analysis of the chemical
residue in plants may be a highly effective method of
assessing the impacts of a site. The bioaccumulative
potential of plants varies greatly however, among
contaminants, contaminant species, soil/sediment texture
and chemistry, plant condition, and genetic composition
of the plant In addition to this variability, plants can
translocate specific contaminants to different pans of the
plant. For example, one contaminant may tend to
accumulate in the roots of a plant whereas a second
contaminant may tend to accumulate in the fruit of the
same plant In this scenario, the collection and analysis
of a plant part that normally does not receive translocated
materials would not result in a useful sample. Therefore.
it is crucial to conduct a literature review prior to
establishing a sampling protocol.
Sampling of herbaceous plants should be conducted
during the growing season of the species of interest
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Sampling of woody plants may be conducted during the
growing or dormant season, however, most plants
translocate materials from the aboveground portions of
the plant to the roots prior to dormancy.
Collection methods and sampling specifics may be found
in U.S. EPA/ERT SOP #2037, Terrestrial Plant
Community Sampling; others are provided in Table 4.1.
3.1.2 Sample Handling and
Preparation
The animals or plants collected should be identified to
species level or the lowest practical taxonomic level.
Appropriate metrics (e.g., weight, animal body length,
plant height) and the presence of any external anomalies,
parasites, and external pathologies should be recorded.
If compositing of the sample material is necessary, it
should be performed in accordance with the study design.
Depending upon the study objectives, it may be necessary
to isolate the contaminant levels in animal tissue from the
contaminant levels in the food or abiotic matrices (e.g.,
sediment) entrained in the digestive tract of the organism.
This is an important process in that it separates the
contribution of two distinct sources of contaminants to the
next trophic level, thereby allowing the data user to
recognize the relative importance of the two sources.
Clearing of the digestive tract (i.e., depuration) of the
organism must then be accomplished prior to the
chemical analysis. The specific depuration procedures
will vary with each type of organism but all involve
allowing the organism to excrete waste products in a
manner in which the products may not be reingested,
absorbed, or deposited back onto the organism.
Biological samples should be handled with caution to
avoid personal injury, exposure to disease, parasites, or
sample contamination. Personal protection such as
gloves should be worn when handling animals and traps
to reduce the transfer of scents or oils from the hand to
the trap, which could cause an avoidance reaction in the
targeted animals.
Samples collected for biological evaluation must be
treated in the same manner as abiotic samples (i.e., the
same health and safety guidelines, decontamination
protocols, and procedures for preventing cross-
contamination must be adhered to). Biological samples
do require some extra caution in handling to avoid
personal injury and exposure to disease, parasites, and
venoms/resins. The selection of sample containers and
storage conditions (e.g., wet ice) should follow the same
protocols as abiotic samples. Refer to Chapter 4.0 for
determination of holding times and additional quality
assurance/quality control (QA/QC) handling procedures.
3.1.3 Analytical Methods
Chemical analytical methods for tissue analysis are
similar to those for abiotic matrices (e.g., soil and water),
however, the required sample preparation procedures
(e.g., homogenization and subsampling) of biological
samples are frequently problematic. For example, large
bones, abundant hair, or high cellulose fiber content may
result in difficult homogenization of mammals and plants.
Extra steps may be required during sample cleanup due
to high lipid (fat) levels in animals tissue or high resin -
content in plant tissue.
Most tissue samples can be placed in a laboratory blender
with dry ice and homogenized at high speeds. The
sample material is then left to sit to allow for the
sublimation of the dry ice. Aliquots of the homogenate
may then be removed for the required analyses.
The requirement for split samples or other QA samples
must be determined prior to sampling to ensure a
sufficient volume of sample is collected. Chapter 4.0
discusses the selection and use of QA/QC samples.
The detection limits of the analytical parameters should
be established prior to the collection of samples.
Detection limits are selected based on the level of
analytical resolution that is needed to interpret the data
against the study objectives. For example, if the detection
limit for a compound is 10 mg/kg but the concentration in
tissue which causes effects is 1 mg/kg, the detection limit
is not artrqiiatr to determine if a problem exists. It should
be noted that standard laboratory detection limits for
abiotic matrices are often not adequate for tissue samples.
Chapter 4.0 provides details on detection limits and other
QA/QC parameters.
The tissue analysis can consist of whole body residue
analysis or analysis of specific tissues (i.e.. fish fillets).
Although less frequently used in Superfund, tissues such
as organs (e.g.. kidney or liver) may be analyzed. The
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study endpoints will determine whether whole body,
fillet, or specific organ samples are to be analyzed.
Concurrent analyses should include a determination of
percent lipids and percent moisture. Percent lipids may
be used to normalize the concentration of non-polar
organic contaminant data. In addition, the lipid content
of the organisms analyzed can be used to evaluate the
organism's health. Percent moisture determinations
allow the expression of contaminant levels on the basis of
wet or dry weight Wet weight concentration data are
frequently used in food chain accumulation models, and
dry weight basis data are frequently reported between
sample location comparisons.
Histopatholoyieal Analysis
Histopathological analysis can be an effective mechanism
for establishing causative relationships due to
contaminants since some contaminants can cause distinct
pathological effects. For example, cadmium causes
visible kidney damage providing causal links between
contaminants and effects. These analyses may be
performed on organisms collected for residue analysis. A
partial necropsy performed on the animal tissue may
indicate the presence of internal abnormalities or
parasites. The time frame and objectives of the study
determine if histopathological analysis is warranted.
3.2 POPULATION/COMMUNITY
RESPONSE STUDIES
Population/community response studies are a commonly
utilized field assessment approach. The decision to
conduct a population/community response study is based
on the rype(s) of contaminants, the time available to
conduct the study, the type of communities potentially
present at the site, and the time of year of the study.
These studies are most commonly conducted on non-
ume-criocal or long-term remediation-type site activities.
During limited time frame responses, however, a
population/community survey or screening level study
may be useful for providing information about potential
impacts associated with a site.
3.2.1 Terrestrial Vertebrate Surveys
Methods for determining adverse effects on terrestrial
vertebrate communities are as follows: censusing or
population estimates, sex-age ratio determinations.
natality/mortality estimations, and diversity studies.
True or accurate censuses are usually not feasible for
most terrestrial vertebrate populations due to logistical
difficulties. Estimations can be derived by counting a
subset of organisms or counting and evaluating signs
such as burrows, nests, tracks, feces, and carcasses.
Capture-recapture studies may be used to estimate
population size but are labor-intensive and usually
require multiple-season sampling. If conducted
improperly, methods for marking captured organisms
may cause irritation or injury or interfere with the
species' normal activities.
Age ratios provide information on natality and rearing
success, age-specific reproductive rates, and mortality
and survival rates. Sex ratios indicate whether sexes are
present in sufficient numbers and proportions for normal
reproductive activity.
Community composition (or diversity) can be assessed by
species frequency, species per unit area, spatial
distribution of individuals, and numerical abundance of
species (Hair 1980).
3.2.2 Benthic Macroinvertebrate
Surveys
Benthic macroinvertebnte (BMI) population/community
evaluations in small- to medium- sized streams have been
successfully used for approximately 100 years to
document injury to the aquatic systems. There are many
advantages to using BMI populations to determine the
potential ecological impact associated with a site.
Sampling is relatively easy, and equipment requirements
are minimal. An evaluation of the community structure
may be used to assess overall water quality, evaluate the
integrity of watersheds, or suggest the presence of an
influence of the community structure that is independent
of water quality and habitat conditions.
Because BMIs are a primary food source for many fish
and other organisms, threats beyond the benthic
community can be inferred from the evaluation of BMIs.
Techniques such as rapid bioassessmem protocols may
be used as a tool to support this type of finding and
inference. A more comprehensive discussion of general
benthological surveys may be found in U.S. EPA (1990).
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3.2.2.1 Rapid Bioassessment Protocols
for Benthic Communities
Rapid bioassessment protocols are an inexpensive
screening tool used for determining if a stream is
supporting or not supporting a designated aquatic life
use. The rapid bioassessment protocols advocate an
integrated assessment, comparing habitat and biological
measures with empirically defined reference conditions
(U.S.EPA1989a).
•\
The three major components of a rapid bioassessment
essential for determining ecological impact are:
• Biological survey
• Habitat assessment
• Physical and chemical measurements
As with all population/community evaluations, the habitat
.assessment is of particular concern with respect to
representative sampling. Care must be taken to prevent
bias during collection of the benthic community resulting
from sampling dissimilar habitats. Similar habitats must
be sampled to make valid comparisons between
locations. In addition to habitat similarity, the sampling
technique and level of effort at each location must be
uniform to achieve an accurate interpretation of results.
In the U.S. EPA Rapid Bioassessment Protocol (RBP),
various components of the community and habitat are
evaluated, a numerical score is calculated, and the score
is compared to predetermined values. A review of the
scores, together with habitat assessment and the physical
and chemical data, support a determination of impact.
U.S. EPA Reference (May. 1989a) presents the
calculation and interpretation of scores!
Standard protocols, including the RBP. have been
developed to facilitate surveying BMls to determine
impact rapidly. These protocols use a standard approach
to reduce the amount of time spent collecting and
analyzing samples. Protocols range from a quick survey
of the benthos' (Protocol I) to a detailed laboratory
classification analysis (Protocol HI). Protocol I may be
conducted in several hours; Protocol n is more intensive
and focuses on major .taxonomic levels; and Protocol IH
may require numerous hours to process each sample to a
greater level of taxonomic and community assessment
resolution. These protocols are used to determine
community health and biological condition via tolerance
values and matrices. They also create and amend a
historical data base that can be used for future site
evaluation.
3.2.22 General Benthological Surveys
Benthological surveys can be conducted with methods
other man those discussed in the RBP protocols utilizing
techniques discussed in the literature. The overall
concept is generally the same as that used in the RBP, but
the specific sampling technique changes depending on
the habitat or community sampled
3.2.2.3 Reference Stations
The use of a reference station is essential to determine
population/community effects attributable to a site. The
use of a reference station within the study area is
preferable (upstream or at a nearby location otherwise
outside the area of site influence). In some cases this is
not possible due to regional impacts, area-wide habitat
degradation, or lack of a similar habitat In these cases
the use of population/community studies should be re-
evaluated within the context of the site investigation. If
the choice is made to include the population/community
study, regional reference or a literature-based evaluation
of the community may be options.
3.2.2.4 Equipment for Benthic Surveys
The selection of the most appropriate sampling
equipment for a particular site is based primarily on the
habitat being sampled. This subsection is a brief
overview of the equipment available for the collection of
BNOs. Detailed procedures are not discussed in this
document For additional information, refer to the SOPs
and methods manuals provided in Table 4.1, or consult
an ecologist/biologist experienced in this type of field
collection.
Long-handled nets or a Surber sampler with a 0.5-
miliimeter (mm) size mesh are common sampling nets for
the collection of macroinvenebrates from a riffle area of
a stream. Samples to be collected from deep water
gravel, sand, or soft boaom habitats such as ponds, lakes,
or rivers are more often sampled using a small Ponar or
Ekrnan dredge. Artificial substrates are used in varying
habitats when habitat matching is problematic and/or
native substrate sampling would not be effective. The
most common types of artificial substrate samplers are
16
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multiple-plate samplers or barbecue basket samplers.
The organisms to be taken to the laboratory for
identification or retained for archival purposes may be
placed in wide-mouthed plastic or glass jars (for ease in
removing contents) and preserved in 70 percent 2-
propanol (isopropyl alcohol) or ethyl alcohol (ethanol),
30 percent formalin, or Kahle's solution. Refer to
methods manuals for detailed information on sample
handling and preservation.
3.2.3 Fish Biosurveys
3.2.3.1 Rapid Bioassessment Protocols
for Fish Biosurveys
RBPs IV and V are two levels of fish biosurvey analyses.
Protocol IV consists of a questionnaire to be completed
with the aid of local and state fisheries experts. Protocol
V is a rigorous analysis of the fish community through
careful species collection, identification, and
enumeration. This level is comparable to the
macroinvenebrate Protocol ID (see Section 3.2.2.1) in
effort Detailed information on both protocols can be
found in Rapid Bioassessments Protocols for Use In
Streams and Rivers (U.S. EPA 1989a).
3.3 TOXICITY TESTS
Toxiciry tests evaluate the relative threat of exposure to
contaminated media (e.g.. soil, sediment, water) in a
controlled setting. These tests are most often conducted
in the laboratory, although they may be conducted in the
field as well. These tests provide an estimate of the
relationship between the contaminated medium, the level
of contaminant and the severity of adverse effects under
specific lest parameters. Toxicity tests are categorized by
several parameters which include duration of the test test
species, life stage of the organism, test end points, and
other variables.
The collection of the actual samples on which the tests
are to be conducted follow the same protocols as
collection of representative samples for chemical
analyses. Typically, a subsample of the media collected
for toxicity testing is submitted for chemical analyses.
The use of a concentration gradient for toxicity testing is
frequently desired to establish a concentration gradient
within the test. This also eliminates the need to sample
all the locations at a site. The specific methods to be
followed for toxicity tests are described in detail in U.S.
EPA's Compendium of ERT Toxiciry Testing
Procedures, OSWER Directive 9360.4-08, EPA/540/P-
91-009 (U.S. EPA 1991a), as well as existing SOPs
listed in Table 4.1. These published procedures address
sample preservation, handling and storage, equipment
and apparatus, reagents, test procedures, calculations,
QA/QC. and data validation. The practical uses of
various toxicity tests, including examples of acute and
chronic tests, are described next Each section includes
an example toxicity test
3.3.1 Examples Of Acute Toxicity
Tests
Example No. 1 (solid-phase soi1>
Laboratory-raised earthworms are placed 30 per replicate
into test chambers containing site soil. A laboratory
control and a site reference treatment are established to
provide a means for comparison of the resulting data set
Depending on the anticipated contaminant concentrations
in the site soil, the soil may be used in its entirety or
diluted with control or site reference soil. The test
chambers are examined daily for an exposure period of
14 days and the number dead organisms is tabulated.
When the observed mortality in the site soil treatments is
statistically compared to control and site reference
treatments, inferences regarding the toxicity of the
contaminant concentrations in the site soil treatments may
be drawn.
Example No. 2 (surface watert
Fathead minnows (Pimephales promelas) are exposed
for 96 hours in aerated test vessels containing surface
water from sampling locations representing a
concentration gradient. The mortality of the organisms is
recorded at the end of the exposure period and
statistically compared to control and site reference
treatments. Statistically significant differences between
treatments may be attributed to the varying contaminant
concentrations.
17
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3.3.2 Examples of Chronic Toxicfty
Tests
Example No. 1 (surface water)
Fathead minnow larvae (Pimcphales promelas) are
exposed for 7 days to surface water collected from
sampling locations that represent a concentration
gradient. Each replicate consists of 20 individuals of the
same maturity level. The test vessels are aerated and the
water is replaced daily. The fish, which should have
remained alive throughout the exposure period, are
harvested and measured for body length and body weight
These results represent growth rates and are statistically
compared to the control and site reference treatments to
infer the lexicological effects of the contaminant
concentrations.
Example No. 2 (sediment^
Midge (Chironomus sp.) larvae are exposed for 10 days
to sediment, overlain with site reference water, and
collected from sampling locations that represent a
concentration gradient Each replicate consists of 200
individuals of the same maturity level (1st instar). The
test vessels are aerated and the water is replaced daily.
At the end of the exposure period, the larvae are removed
from the test vessels and measured for body length and
body weight
The organisms are then returned to the test vessels and
allowed to mature to the adult stage. An emergence trap
is placed over the test vessel and the number of emerging
adults is recorded. These results, as well as the length
and weight results, are statistically compared to the
control and site reference treatments to infer the
lexicological effects of the contaminant concentrations.
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Figure 2: Common Mammal Traps
Havahart Trap
Longworth live trap
(A)
(B)
Folding (A) and non-folding (B) Sherman live traps
19
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TABLE I
Reference Lisl of Standard Operating Procedures -: Ecological Sampling Methods
SOP/Method No.
SOP No 1820
SOP No 1821
SOP No 1822
SOP No 1823
SOP No 2020
SOP No. 2021
SOP No 2022
SOP No 2023
SOP No. 2024
SOP No. 2025
SOP No 2026
SOP No. 2027
SOP No. 2028
SOP No 1001
SOP No. 1-002
Greene etil.( 1989)
SOP No. 1 005
SOP No. 2029
SOP No. 2032
SOP No. 2033
SOP No. 2034
SOP No. 2035
SOP No. 2036
SOP No 2037
Source
ERTC
ERTC
ERTC
ERTC
ERTC
ERTC
ERTC
ERTC
ERTC
ERTC
ERTC
ERTC
ERTC
ERTC
ERTC
.
ERTC
ERTC
ERTC
ERTC
ERTC
ERTC
ERTC
ERTC
Procedure/Melhod Title
Tissue Homogenizalion Procedure
Semi-Volaliles Analysis of Tissue Samples by GC/MS
Peslicides/ICB Analysis of Tissue Samples by GC/ECD
Microwave Digestion and Metals Analysis of Tissue Samples
7-Day Standard Reference Toxicily Test Using Larval Fathead Minnows Pimephalts promelat
24 Hour Range Finding Test Using Daphnia magnaot Daphnia pule*
96-Hour Acute Toxicily Test Using Larval fimephalet promelai
24-Hour Range Finding Test Using Larval PimepHalei promelai
48-Hour Acute Toxicity Test Using Daphnia magna or Daphnia pule*
7-Day Renewal Toxicily Test Using Ceriodaphnia dubia
7-Day Static Toxicily Test Using Larval FimepHalei promtlas
96- Hour Static Toxicity Test Using Stlenailrum capricomulum
10- Day Chronic Toxicity Test Using Daphnia magna'ot Daphnia pulex
15-Day Solid Phase Toxicily Test Using Chironomus lenlani
28-Day Solid Phase Toxicily Test Using Hyalella atleca
14-Day Acute Toxicily Test Using adult Eiienia andrei (earthworms)
Field Processing of Fish
Small Mammal Sampling and Processing
Benthic Sampling
Plant Protein Determination
Plant Biomass Determination
Plant Peroxidase Activity Determination
Tree Coring and Interpretation
Terrestrial Plant Community Sampling
Publication No
(In development)
(in development)
(in development)
(in development)
OSWER EPA/540/P 91/009
OSWER EPA/540/P-9 1/009
OSWER EPA/540/P-9 1/009
OSWER EPA/54(VP-9i/009
OSWER EPA/540/P-9I/009
OSWER EPA/540/P-9 1/009
OSWER EPA/540/P-9I/009
OSWER EPA/540/P-9I/009
OSWER EPA/540/P-9 1/009
(in development)
(in development)
EPA 600/3-88-029
(In development)
(in development)
(in development)
(In development)
(In development)
(In development)
(In development)
(in development)
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4.0 QUALITY ASSURANCE/QUALITY CONTROL
4.1 INTRODUCTION
The goal of representative sampling is to yield
quantitative data that accurately depict site conditions in
a given period of time. QA/QC measures specified in the
sampling procedures minimize and quantify the error
introduced into the data.
Many QA/QC measures are dependant on QA/QC
samples submitted with regular field samples. QA/QC
samples evaluate the three following types of information:
(1) the degree of site variation; (2) whether samples were
cross-contaminated during sampling and sample handling
procedures; and (3) whether a discrepancy in sample
results is attributable to field handling, laboratory
handling, or analysis. For additional information on QA
•objectives, refer to U.S. EPA Quality Assurance/Quality
Control (QA/QC) Guidance for Removal Activities,
EPA/540/G-90/004. April 1990.
4.2 DATA CATEGORIES
The U.S. EPA has established a process of data quality
objectives (DQOs) which establish what type, quantity.
and quality of environmental data are appropriate for
their intended application. In its DQO process, U.S.
EPA has denned two broad categories of data: screening
and definitive.
Screening data are generated by rapid, less precise
methods of analysis with less rigorous sample
preparation. Sample preparation steps may be restricted
to simple procedures such as dilution with a solvent,
rather than an elaborate extraction/digestion and cleanup.
At least 10 percent of the screening data are confirmed
using the analytical methods and QA/QC procedures and
criteria associated with definitive data. Screening data
without associated confirmation data are not considered
to be data of known quality. To be acceptable, screening
data must include the following:
• chain of custody
• initial and continuing calibration
• analyte identification
• analyte quantification
Streamlined QC requirements are the defining
characteristic of screening data.
Definitive data are generated using rigorous analytical
methods (e.g., approved U.S. EPA reference methods).
These data are analyte-specific, with confirmation of
analyte identity and concentration. Methods produce
tangible raw data (e.g., chromatograms, spectra, digital
values) in the form of hard-copy printouts or computer-
generated electronic files. Data may be generated at the
site or at an off-site location as long as the QA/QC
requirements are satisfied. For the data to be definitive,
either analytical or total measurement error must be
determined. QC measures for definitive data contain all
the elements associated with screening data, but also
include trip, method, and rinsate blanks; matrix spikes;
performance evaluation samples; and replicate analyses
for error determination.
For more details on these data categories, refer to U.S.
EPA Data Quality Objectives Process For Superfund,
EPA/540/R-93/071, Sept 1993.
4.3 SOURCES OF ERROR
The four most common potential sources of data error in
biological sampling:
• Sampling design
• Sampling methodology
• Sample heterogeneity
• Sample analysis
4.3.1 Sampling Design
The initial selection of a habitat is a potential source of
bias in biological sampling, which might either
exaggerate or mask the effects of hazardous substances in
the environment. In a representative sampling scheme,
habitat characteristics such as plant and animal species
composition, substrates, and degree of shading should be
similar at all locations, including the reference location.
The same individual should select both the test site and
the control and background site to minimize error in
comparing site conditions.
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Standardized procedures for habitat assessment and
selection also help minimize design error. The selection
of an inappropriate species may introduce an error into
the representative sampling design. This error can be
minimized by selecting a species that is representative of
the habitat and whose life-cycle is compatible with the
timing of the study. In addition, migratory or transient
species should be avoided.
4.3.2 Sampling Methodology
Sampling methodology and sample handling procedures
may contain possible sources of error such as unclean
sample containers, improper sample handling, and
improper shipment procedures. Procedures for sample
collection and handling should be standardized to allow
easier identification of potential error. Follow SOPs or
established procedures to ensure that all sampling
techniques are performed consistently despite different
. sampling teams, dates, or locations. Use QA/QC
samples (Section 4.4) to evaluate errors due to improper
sampling methodology and sample handling procedures.
These guidelines should apply to biological as well as
soil, sediment, and water sampling.
During fishing operations, the sampling crew can prevent
habitat disturbance by staying out of the water body near
the sampling locations. The use of any particular
technique may introduce judgment error into the
sampling regimen if done improperly. For all techniques,
sampling should be conducted from the downstream
location to the upstream location to avoid contamination
of the upstream stations. Data comparability is
maintained by using similar collection methods and
sampling efforts at all stations.
Rapid bioassessments in the field should include two
QA/QC procedures: 1) collection of replicate samples at
stauons to check on the accuracy of the collection effort, -
and 2) repeal a portion (typically 10%) recount and
^identification for accuracy. ,
For tissue analyses, tools and other sampling equipment
should be dedicated to each sample, or must be
decontaminated between uses. To avoid contamination,
sample containers must be compatible with the intended
tissue matrix and analysis.
4.3.3 Sample Heterogeneity
Tissues destined for chemical analysis should be
homogenized Ideally, tissue sample homogenates should
consist of organisms of the same species, sex, and
development stage and size since these variables all affect
chemical uptake. There is no universal SOP for tissue
homogenization; specific procedures depend on the size
and type of the organism. For example, tissues must be
cut from fur and shell-bearing organisms as they cannot
be practically homogenized as a whole. Homogenization
procedures may vary by site objective. Tissue
homogenates should be stored away from light and kept
frozen at -20° C. Tissue homogenates are prepared in
the laboratory and could be subject to cross-
contamination.
Refer to U.S. EPA/ERT SOP #1820, Tissue
Homogenizaaon Procedures for further details on tissue
homogenization procedures.
4.3.4 Sample Analysis
Analytical procedures may introduce errors from
laboratory cross-contamination, extraction difficulties,
and inappropriate methodology. Fats naturally present in
tissues may interfere with sample analysis or extraction
and elevate detection limits. Detection limits in the tissue
samples must be the same as in the background tissue
samples if a meaningful comparison is to be made. To
minimize this interference, select an extraction or
digestion procedure applicable to tissue samples.
Because many compounds (e.g., chlorinated
hydrocarbons) concentrate in fatty tissues, a percent lipid
analysis is necessary to normalize results among samples.
Lipid recoveries vary among different analytical methods;
percent lipid results for samples to be normalized and
compared must be generated by the same analytical
method. Select a lipid analysis based on the objective of
the study (see references Heroes and Allen [1983] and
Bligh and Dyer 1959). Sample results may be
normalized on a wet-weight basis. If sample results are
to be reported on a dry-weight basis, instruct the
analytical laboratory to report the percent moisture
content for each sample.
Appropriate sample preservation prevents loss of
compounds and decomposition of tissues before analysis.
Consult the appropriate SOP, analytical method, or
designated laboratory contact to confirm holding times for
tissue samples.
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Tissue samples destined for sorting and identification
(e.g,, benthic macroinvenebrates. voucher fish) should be
preserved in isopropyl or ethyl alcohol, formalin, or
Kahle's solution. Preservation in these solvents precludes
any chemical analysis.
4.4 QA/QC SAMPLES
QA/QC samples are collected at the site as prepared by
the laboratory. Analysis of the QA/QC samples provides
information on the variability and usability of biological
sampling data, indicates possible field sampling or
laboratory error, and provides a basis for future validation
and usability of the analytical data. The most common
field QA/QC samples are field replicates, reference, and
rinsate blank samples. The most common laboratory
QA/QC samples are performance evaluation (PE), matrix
spike (MS), and matrix spike duplicate (MSD) samples.
QA/QC results may suggest the need for modifying
sample collection, preparation, handling, or analytical
procedures if the resultant data do not meet site-specific
quality assurance objectives.
Refer to data validation procedures in U.S. EPA Quality
Assurance/Quality Control (QA/QC) Guidance for
Removal Activities, EPA/540/G-9Q/004. April 1990, for
guidelines on utilizing QA/QC samples.
4.4.1 Replicate Samples
Field Replicates
Field replicates for solid media are samples obtained
from one sampling point that are homogenized, divided
into separate containers, and treated as separate samples
throughout the remaining sample handling and analytical
processes. Field replicates for aqueous samples are
samples obtained from one location that are homogenized
and divided into separate containers. There are no "true"
field replicates for biological samples, however,
biological samples collected from the same station are
typically referred to as replicates. In this case, the
biological replicates are used to determine the variability
associated with heterogeneity within a biological
population. Field replicates may be sent to two or more
laboratories or to the same laboratory as unique samples.
Field replicates may be used to determine total error for
critical samples with contaminant concentrations near the
level that determines environmental impact To
determine error, a minimum of eight replicate samples is
recommended for valid sgtjyrirai analysis.. For total error
determination, samples should be analyzed by the same
laboratory. The higher detection limit associated with
composite samples may limit the usefulness of error
determination.
NOTE: A replicate biological sample may consist of
more than a single organism in those cases where the
species mass is less than the mass required by the
analytical procedure to attain required detection limits.
This variability in replicate biological samples is
independent of the variability in analytical procedures.
Toxicitv Testing Replicates
For sediment samples, at least 3 replicate treatments
should be conducted to determine variability between
testsJThe function of these replicates is to determine the-
variability of the test organism population within each
treatment This assumes the sample matrix exhibits a
uniform concentration of the contaminants of concern
within each treatment Large variability may indicate a
problem with the test procedures or organisms or lack of
contaminant homogeneiry{within the sample matrix.
Site-Sr-Tic Fj
Examle No. 1
of the Use of Replicates
Two contaminant sources were identified at an active
copper smelting facility. The first area was a slag pile
containing high levels of copper suspected of migrating
into, the surrounding surface runoff pathways,
subsequently leaching into the surface water of a
surrounding stream system. The second area was the
contaminated creek sediment that was present in the
drainage pathway of the slag pile.
Whole-phase sediment toxicity tests were selected to
evaluate the toxicity associated with the copper levels in
the stream sediments. Sediment was collected at each
sampling location (six locations total) to provide the
testing laboratory with sufficient sample volume to
perform these evaluations. Ten-day static renewal tests
using the amphipod, Hyalella azteca, and the midge,
Chironomus lentans, were chosen. The toxicity test
utilized four "replicates" per sampling location (or
treatment), each replicate containing fifteen organisms.
The purpose of these replicates was to determine the
23
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variability within the test organism population within
each treatment.
The results reported mean survival for Hyalella azteca in
the contaminated sediment (8 to SO percent) to be
significantly lower than survival in the uncontaminated
reference sediment (85 percent). Similarly, mean
survival for Chironomus unions in the contaminated
sediment (0 to 63 percent) was significantly lower than
survival in the uncontaminated reference sediment (83
percent).
Example T^p, ^
An inactive manufacturing facility had stored its stock
compounds in unprotected piles for a number of years,
resulting .in DDT contamination of the adjacent
watershed. DDT contamination in a stream located
adjacent to the site extended from the manufacturing
facility to approximately 27 miles downstream.
A field study was designed to quantitatively determine if
the levels of DDT in the water and sediment in this
stream were resulting in an adverse ecological impact.
This was accomplished through the examination of
several in situ environmental variables in conjunction
with laboratory analyses. Water, sediment, and resident
biota were collected and submitted for various physical
and chemical determinations. Additional sediments were
secured and utilized for toxicity testing with three
surrogate species. Finally, the benthic invertebrate
community was sampled and the structure and function of
this segment of the aquatic ecosystem evaluated.
Benthic invertebrates were collected from three areas at
each sampling location (i.e., three "replicates" per
location) and evaluated for various quantitative
community metrics. The purpose of these replicates were
to determine the spatial variability in the stream among
the three areas within each sampling location.
Community structure, diversity indices, taxonomic
evenness, an evaluation of the function feeding groups.
and statistical analyses were performed on the data set.
Qualitative and statistical comparison of the results
between the contaminated areas and the uncontaminated
reference indicated that the benthic invertebrate
community was adversely affected downstream of the site
compared to the upstream reference. Taxonomic and
functional diversity varied inversely with DDT levels in
sediment and water. These results were further
substantiated by the toxicity evaluation results.
Example No. 3
Phase I and D Remedial Investigation and Feasibility
Studies (RTFS) have indicated that the soils surrounding
an industrial and municipal waste disposal site were
contaminated with PCBs. A preliminary site survey
revealed the presence of small mammal habitat and
mammal signs in the natural areas adjacent to the site as
well as an area that appeared to be outside-of the site's
influence (ix.. a potential reference area). A site
investigation was subsequently conducted to determine
the levels of PCBs accumulating into the resident
mammal community from contact with the PCB-
contaminated soil.
Three small mammal trapping areas were identified for
this site. Two areas were located in PCB-contanunated
areas, the third area was a reference. Trapping grids .
were established in each area consisting of 100 traps of
various design. Six soil samples were also collected from
each trapping area to characterize the levels of PCBs
associated with the anticipated captured mammals.
A total of 32 mammals were collected at this site.
Twelve were collected from each on-site area and six
were collected from the reference area. All captured
mammals were submitted for whole body analysis of
PCBs. Mean PCS concentrations in the mammals were
as follows: on-site areas (1250 and 1340 wg/kg, wet
weight); reference area (490 yg/kg, wet weight). A one-
way analysis of variance was conducted on the data set
treating each animal in an area as a "replicate" (i.e.. 12
replicates from each on-site area and 6 replicates from
the reference). The results of the statistical analyses
indicated that there was a statistically significant
difference between on-site and reference area PCB levels
in the mammals (p<0.lO). Therefore, in this example.
there were no analytical replicates since each individual
mammal was analyzed. However, each mammal
represented a statistical replicate within each trapping
area.
4.4.2 Collocated Samples
A collocated sample is collected from an area adjoining
a field sample to determine variability of the matrix and
contaminants within a small area of the site. For
example, collocated samples for chemistry analysis split
24
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from the sample collected for the toxicity test are
collected about one-half to three feet away from the field
sample location. Plants collected from within the same
sampling plot may be considered collocated. Collocated
samples are appropriate for assessing variability only in
a small area, and should not be used to assess variability
across the entire site or for assessing error.
4.4.3 Reference Samples
Reference biological samples may be taken from a
reference area outside the influence of the site.
Comparison of results from actual samples and samples
from the reference area may indicate uptake, body
burden, or accumulation of chemicals on the site. The
reference area should be close to the site. It should have
habitats, size and terrain similar to the site under
investigation. The reference site need not be pristine.
Biological reference samples should be of the same
.species, sex, and developmental stage as the field site
sample.
4.4.4 Rinsate Blank Samples
A nnsate blank is used to assess cross-contamination
from improper equipment decontamination procedures.
Rinsate blanks are samples obtained by running analyte-
frec water over decontaminated sampling equipment.
Any residual contamination should appear in the rihsate
data. Analyze the rinsate blank for the same analytical
parameters as the field samples collected that day. When
dedicated cutting tools or other sampling equipment are
not used, collect one rinsate blank per device per day.
4.4.5 Field Blank Samples
Field blanks are samples prepared in the field using
certified clean water or sand that are then submitted to the
laboratory' for analysis. A field blank is used to evaluate
contamination or error associated with sampling
methodology, preservation, handling/shipping, and
laboratory procedures. If appropriate for the test, submit
one field blank per day.
4.4.6 Trip Blank Samples
Trip blanks are samples prepared prior to going into the
field. They consist of certified clean water or sand, and
they are not opened until they reach the laboratory. Use
tnp blanks when samples are being analyzed for volatile
organics. Handle, transport, and analyze trip blanks-in
the same manner as the other volatile organic samples
collected that day. Trip blanks are used to evaluate error
associated with sampling methodology, shipping and
handling, and analytical procedures, since any volatile
organic contamination of a trip blank would have to be
introduced during one of those procedures.
4.4.7 Performance Evaluation
/Laboratory Control Samples
A performance evaluation (PE) sample evaluates the
overall error from the analytical laboratory and detects
any bias in the analytical method being used. PE samples
contain known quantities of target analytes manufactured
under strict quality control. They are usually prepared by
a third party under a U.S. EPA certification program.
The samples are usually submitted "blind" to analytical
laboratories (the sampling team knows the contents of the
samples, but the laboratory does not). Laboratory
analytical error (usually bias) may be evaluated by the
percent recoveries and correct identification of the
components in the PE sample.
4.4.8 Controls
Analytical Laboratory Control Samples
A chemical analytical laboratory control sample (LCS)
contains quantities of target analytes known to the
laboratory and are used to monitor "controlled"
conditions. LCSs are analyzed under the same sample
preparation, reagents, and analytical methods as the field
samples. LCS results can show bias and/or variability in
analytical results.
ToKicirv Testing Control Groups
In toxicity tests, a laboratory reference toxicant treatment
and a control treatment are both typically utilized in
addition to a site reference treatment. This test involves
exposing the test organism population to a standardized
reference toxicant at a standardized dose, then comparing
the response to historical laboratory records for that
culture. The mortality results of the newly conducted
reference toxicant test should be similar to the historical
results. This is conducted to reveal if the generation(s) in
the present culture is viable for use in the toxicity test, or
if the culture has grown resistant or intolerant to the
toxicant over time. Therefore, a laboratory reference
25
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toxicant test should be conducted prior to the testing of
the site matrices.
In contrast, a laboratory control test is conducted
simultaneously with the testing of the site matrices. This
treatment identifies mortality factors that are unrelated to
site contaminants. This is accomplished by exposing the
test organism population to a clean dilution water and/or
a clean laboratory substrate.
4.4.9 Matrix Spike/Matrix Spike
Duplicate Samples /
Matrix spike and matrix spike duplicate samples
(MS/MSDs) are supplemental volumes of field-collected
samples that are spiked in the laboratory with a known
concentration of a target analyte to determine matrix
interference. Matrix interference is determined as a
function of the percent analyte recovery in the sample
extraction. The percent recovery from MS/MSDs
indicates the degree to which matrix interferences will
affect the identification and/or quantitation of a substance.
MS/MSDs can also be used to monitor laboratory
performance. When two or more pairs of MS/MSDs are
analyzed, the data obtained may also be used to evaluate
error due to laboratory bias and precision. Analyze one
MS/MS D pair to assess bias for every 10 samples, and
use the average percent recovery for the pair. To assess
precision, analyze at least eight matrix spike replicates
from the same sample, and determine the standard
deviation and the coefficient of variation. See the U.S.
EPA Quality Assurance/ Quality Control (QA/QC)
Guidance for Removal Activities (April 1990) for
directions on calculating analytical error.
MS/MSDs are a required QA/QC element of the
definitive data objectives. MS/MSDs should accompany
every 10 samples. Since the MS/MSDs are spiked field
samples, sufficient volume for three separate analyses
must be provided. Organic analysis of tissue samples is
frequently subject to matrix interferences which causes
biased analytical results. Matrix spike recoveries are
often low or show poor precision in tissue samples. The
matrix interferences will be evident in the matrix spike
results. Although metals analysis of tissue samples is
usually not subject to these interferences. MS/MSD
samples should be utilized to monitor method and
laboratory performance. Some analytical parameters
such as percent hpids. organic carbon, and panicle-size
distribution are exempt from MS/MSD analyses.
4.4.10 Laboratory Duplicate
Samples
A laboratory duplicate is -a sample that undergoes
preparation and analysis twice. The laboratory takes two
aliquots of one sample and treats them as if they were
separate samples. Comparison of data from the two
analyses provides a measure of analytical reproducibility
within a sample set. Discrepancies in duplicate analyses
may indicate poor homogenization in the field or other
sample preparation error, whether in the field or in the
laboratory. However, duplicate analyses are not possible
with most tissue samples unless a homogenate of the
sample is created.
4.5 Data Evaluation
4.5.1 Evaluation of Analytical Error
Analytical error becomes significant in decision-making
as sample results approach the level of environmental
impact The acceptable level of error is determined by
the intended use of the data and litigation concerns. To
be definitive, analytical data must have quantitative
measurement of analytical error with PE samples and
replicates. The QA samples identified in this section can
indicate a variety of qualitative and quantitative sampling
errors. Due to matrix interferences, causes of error may
be difficult to determine in organic analysis of tissue
samples.
4.5.2 Data Validation
Data from tissue sample analysis may be validated
according to the Contract Laboratory Program National
Functional Guidelines (U.S. EPA 1994) and according to
US. EPA Quality Assurance/Quality Control (QA/QC)
Guidance for Removal Activities, EPA/540/G-90/004,
April 1990. Validation of organic data may require an
experienced chemist due to complexity of tissue analysis.
26
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5.0 DATA ANALYSIS AND INTERPRETATION
5.1 INTRODUCTION
The main objective of biological surveys conducted at
Supertund sites is the assessment of site-related threat or
effect. For many types of biological data (e.g., levels of
contaminants in organisms collected on site and from a
reference location), hypotheses are tested to determine
the presence or absence of an effect For some biological
tests (e.g., benthic macroinvertebratc studies, toxicity
tests), the data analysis and interpretation process is
outlined in existing documents (U.S. EPA November
1990, U.S. EPA May 1996). For many Superfund
ecological assessments, a weight-of-evidence approach
is used to interpret the results of different studies or tests
conducted at a site.
The statistical tests and methods that will be employed
should be based on the objective of the data evaluation.
These components should be outlined in the Work Plan
or Sampling and Analysis Plan. This process will help
focus the study to ensure that the appropriate type and
number of samples are collected.
5.2 DATA PRESENTATION AND
ANALYSIS
5.2.1 Data Presentation Techniques
In many cases, before descriptive statistics are calculated
from a data set. it is useful to try various graphical
displays of the raw data. The graphical displays help
guide the choice of any necessary transformations of the
data set and the selection of appropriate statistics to
summarize the data. Since most statistical procedures
require summary statistics calculated from a data set. it is
important that the summary statistics represent the entire
data set. For example, the median may be a more
appropriate measure of central tendency than the mean
for a data set that contains outliers. Graphical display of
a data set could indicate the need to log transform data so
that symmetry indicates a normal distribution. Four of the
most useful graphical techniques are described next.
A histogram is a bar graph that displays the distribution
of a data set. and provides information regarding the
location of the center of the sample, amount of dispersion.
extent of symmetry, and existence of outliers. Stem and
leaf plots are similar to histograms in that they provide
information on the distribution of a data set; however they
also contain information on the numeric values in the data
set Box and whisker plots can be used to compare two
or more samples of the same characteristic (e.g., stream
ffil values for two or more years). Scatter plots are a
useful method for examining the relationship between
two sets of variables. Figure 4 illustrates the four graph
techniques described previously.
5.2.2 Descriptive Statistics
Large data sets are often summarized using a few
descriptive statistics. Two important features of a set of
data are the central tendency and the spread. Statistics
used to describe central tendency include the arithmetic
mean, median, mode and geometric mean. Spread or
dispersion in a data set refers to the variability in the
observations about the center of the distribution.
Statistics used to describe data dispersion include range
and standard deviation. Methods for calculating
descriptive statistics can be found in any statistics
textbook, and many software programs are available for
statistical calculations.
5.2.3 Hypothesis Testing
Biological studies are conducted at Superfund sites to
determine adverse effects due to site-related factors. For
many types of biological data, hypothesis testing is the
statistical procedure used to evaluate data. Hypothesis
testing involves statistically evaluating a parameter of
concern, such as the mean or median, at a specified
probability for incorrectly interpreting the analysis
results. In conventional statistical analysis, hypothesis
testing for a trend or effect is based on a null hypothesis.
Typically, the null hypothesis is presumed when there is
no trend or effect present. To test this hypothesis, data
are collected to estimate an effect The data are used to
provide a sample estimate of a test statistic, and a table
for the test statistic is consulted to determine how unlikely
the observed value of the statistic is if the null hypothesis
is true. If the observed value of the test statistic is
unlikely, the null hypothesis is rejected. In ecological risk
assessment a hypothesis is a question about the
relationship among assessment endpoints and their
27
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predicted responses when exposed to contaminants. The
most basic hypothesis that is applicable to virtually all
Superfund sites is that site-related contaminants are
causing adverse effects of the assessment endpoint(s).
5.3 DATA INTERPRETATION
5.3.1 Chemical Residue Studies
Chemical residue data may be evaluated in two ways.
First, the contaminant concentrations by themselves
provide evidence of bioaccumulation and probable food
chain transfer of the contaminants, and an overall picture
of the distribution of contaminants in the biological
community. Second, the residue data may be evaluated
against literature residue values that are known to cause
no effect or an adverse effect in the organism.
5.3.2 Population/Community
Studies
The interpretation of population/community data is
extensive, therefore, the reader is referred to a detailed
treatment in U.S. EPA (November 1990), U.S. EPA
(1989a), Karr et al. (1986). and other literature.
5.3.3 Toxicity Testing
Measurement endpoints obtained in toxicity tests are
.generally compared to results from a laboratory control
and a reference location sample to determine whether
statistically significant differences exist. If significant
effects (e.g.. mortality, decreased reproduction) are
observed, additional statistical analyses can be run to
determine whether observed effects correlate with
measured contaminant levels. The reader is referred to a
detailed treatment in ASTM (1992). U.S. EPA (May
1988). U.S. EPA (March 1989b).
5.3.4 Risk Calculation
Preliminary screening value results are interpreted by
comparison of histohcal and/or new site analytical data
against literature toxicity values. This comparison will
suggest if the probability of risk exists and whether
additional evaluation is desired.
assessment, mathematical models, such as the Hazard
Quotient method, are used to evaluate the site data
against literature toxicity values. Based on the type of
model used, the results can be extrapolated to suggest the
presence of ecological risk.
If the evaluation is pursued to an ecological risk
28
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Figure 3 Illustrations of Sample Plots
IBI DATA
12 25 33 56
12 24 34 SB
14 26 35
15 24 36
16 24 35
22 27 3B
24 23 41
23 28 42
A) Histogram
B) Leaf Plot
250 300 350 '
Sediment Zinc (mg/Kg)
C) Whisker Plot D) Scatter Plot
29
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APPENDIX A - CHECKLIST FOR ECOLOGICAL
ASSESSMENT/SAMPLING
Introduction
The checklist that follows provides guidance in making observations for an ecological assessment It is not intended for
limited or emergency response actions (e.g., removal of a few drums) or for purely industrial settings with no discharges.
The checklist is a screening tool for preliminary site evaluation and may also be useful in planning more extensive site
investigations. It must be completed as thoroughly as time allows. The results of the checklist will serve as a starting point
for the collection of appropriate biological data to be used in developing a response action. It is recognized that certain
questions in this checklist are not universally applicable and that site-specific conditions will influence interpretation.
Therefore, a site synopsis is requested to facilitate final review of the checklist by a mined ecologist
Checklist
The checklist has been divided into sections that correspond to data collection methods and ecosystem types. These sections
are:
I. Site Description
1A. Summary of Observations and Site Setting
IL Terrestrial Habitat Checklist
HA. Wooded
Iffi. Shrub/Scrub
EC. Open Field .
ED. Miscellaneous
m. Aquatic Habitat Checklist - Non-Flowing Systems
IV. Aquatic Habitat Checklist - Flowing Systems
V. Wetlands Habitat Checklist
30
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Checklist for Ecological Assessment/Sampling
I. SITE DESCRIPTION
1. Site Name:.
Location: _
County: City: State:.
2. Latitude: , Longitude:
3. What is the approximate area of the site?.
4. Is this the first site visit? D yes D no If no, attach trip report of previous site visit(s), if available.
Date(s) of previous site visitfs):
5. Please attach to the checklist USGS topographic map(s) of the site, if available.
6. Are aerial or other site photographs available? D yes D no If yes, please attach any available photo(s) to the site
map at the conclusion of this section.
31
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7. The land use on the site is:
% Urban
_% Rural
% Residential
% Industrial (D light D heavy)
% Agricultural
(Crops:
% Recreational
(Describe; note if it is a park, etc.)
_% Undisturbed
_% Other
The area surrounding the site is:
mile radius
% Urban
% Rural
_, % Residential
% Industrial (D light D heavy)
% Agricultural
(Crops: ;
Recreational
(Describe; note if it is a park, etc.)
_% Undisturbed
_% Other
8. Has any movement of soil taken place at the site? D yes D no. If yes, please identify the most likely cause of this
disturbance:
Agricultural Use
Natural Events
Please describe:
. Heavy Equipment
Erosion
. Mining
.Other
32
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9. Do any potentially sensitive environmental anus exist adjacent to or in proximity to the site, e.g., Federal and State
paries. National and State monuments, wetlands, prairie potholes? Remember, flood plains and wetlands are not
always obvious; do not answer "no" without confirming information.
Please provide the source(s) of information used to identify these sensitive areas, and indicate their general location
on the site map.
10. What type of facility is located at the site?
D Chemical O Manufacturing D Mixing D Waste disposal
D Other (specify)
11. What are the suspected contaminants of concern at the site? If known, what are the maximum concentration levels?
12. Check any potential routes of off-site migration of contaminants observed at the site:
O Swales D Depressions D Drainage ditches
" Runoff O Windblown paniculate* D Vehicular traffic
[3 Other (specify)
13. If known, what is the approximate depth to the water table?.
14. Is the direction of surface runoff apparent from site observations? D yes D no If yes, to which of the following
does the surface runoff discharge? Indicate all that apply.
Z Surface water U Groundwater D Sewer O Collection impoundment
IS. Is there a navigable waterbody or tributary to a navigable waterbody? D yes O no
33
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16. Is there a wateibody anywhere on or in the vicinity of the site? If yes, also complete Section ffl: Aquatic Habitat
Checklist - Non-Flowing Systems and/or Section IV: Aquatic Habitat Checklist - Flowing Systems.
D yes (approx. distance ) D no
17. Is there evidence of flooding? OyesDno Wetlands and flood plains are not always obvious; do not answer "no"
without confirming information. If yes, complete Section V: Wetland Habitat Checklist
18. If a field guide was used to aid any of the identifications, please provide a reference. Also, estimate the time spent
identifying fauna. [Use a blank sheet if additional space is needed for text]
19. Are any threatened and/or endangered species (plant or animal) known to inhabit the area of the site? D yes D no
If yes. you are required to verify this information with the U.S. Fish and Wildlife Service. If species' identities are
known, please list them next
20. Record weather conditions at the time this checklist was prepared:
DATE:__ '
Temperature (°C/CF) Normal daily high temperature
Wind (direction/speed) Precipitation (rain, snow)
Cloud cover
34
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1A. SUMMARY OF OBSERVATIONS AND SITE SETTING
Completed by i Affiliation.
Additional Preparers.
Site Manager
Date
35
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IL TERRESTRIAL HABITAT CHECKLIST
DA. WOODED
1. Are there any wooded areas at the site? Dyes D no If no, go to Section HE: Shnib/Scnib.
2. What percentage or area of the site is wooded? ( _% acres). Indicate the wooded area on the site map
which is attached to a copy of this checklist. Please identify what information was used to determine the wooded
area of the site.
3. What is the dominant type of vegetation in the wooded area? (Circle one; Evergreen/Deciduous/ Mixed) Provide a
photograph, if available.
Dominant plant, if known:.
4. What is the predominant size of the trees at the site? Use diameter at breast height
D 0-6 in. D 6-12 in. D>12in.
5. Specify type of understory present, if known. Provide a photograph, if available.
ITB. SHRUB/SCRUB
1. Is shrub/scrub vegetation present at the site? D yes D no If no, go to Section EC: Open Field.
2. What percentage of the site is covered by scrub/shrub vegetation? ( % acres). Indicate the areas of
shrub/scrub on the site map. Please identify what information was used to determine this area.
3. What is the dominant type of scrub/shrub vegetation, if known? Provide a photograph, if available.
4. What is the approximate average height of the scrub/shrub vegetation?
D 0-2 ft. D 2-5 ft. D > 5 ft.
36
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S. Based on site observations, how dense is the scrub/shrub vegetation?
D Dense D Patchy D Sparse
IIC OPEN FIELD
1. Are there open (bare, barren) field areas present at the site? Dyes Ono If yes, please
indicate the type below:
D Prairie/plains D Savannah D Old field D Other (specify).
2. What percentage of the site is open field? ( % ___ acres). Indicate the open fields on the site map.
3. What is/are the dominant plant(s)? Provide a photograph, if available.
4. What is the approximate average height of the dominant plant?.
5. Describe the vegetation cover D Dense D Sparse D Patchy
ITO. MISCELLANEOUS
1. Are other types of terrestrial habitats present at the site, other than woods, scrub/shrub, and open field? Dyes Dno
If yes. identify and describe them below.
2. Describe the terrestrial miscellaneous habitat(s) and identify these area(s) on the site map.
37
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3. What observations, if any, were made at the site regarding the presence and/or absence of insects, fish, birds,
mammals, etc.?
4. Review the questions in Section I to determine if any additional habitat checklists should be completed for this site.
38
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AQUATIC HABITAT CHECKLIST - NON-FLOWING SYSTEMS
Note: Aquatic systems are often associated with wetland habitats. Please refer to Section V, Wetland Habitat
Checklist.
1. What type of open-water, non-flowing system is present at the site?
O Natural (pond, lake)
D Artificially created (lagoon, reservoir, canal, impoundment)
2. If known, what is the iiame(s) of the waterbody(ies) on or adjacent to the site?
3. If a waterbpdy is present, what are its known uses (e.g.: recreation, navigation, etc.)?
*'
4. What is the approximate size of the waterbody(ies)? acre(s).
5. Is any aquatic vegetation present? O yes D no If yes. please identify the type of vegetation present if known.
O Emergent D Submergent O Floating
6. If known, what is the depth of the water?
7. What is the general composition of the substrate? Check all that apply.
Z Bedrock G Sand (coarse) D Muck (fine/black)
Z Boulder (>10 in.) ' D Silt (fine) D Debris
Z Cobble (2.5-10 in.) C Marl (shells) D Detritus
Z Gravel (0.1-2.5 in.) D Clay (slick) D Concrete
Z Other (specify) '
8. What is the source of water in the waterbody?
Z River/Stream/Creek D Groundwater D Other (specify).
Z Industrial discharge D Surface runoff
39
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9. Is there a discharge from the site to the waterbody? Dyes Dno If yes, please describe this
discharge and its path.
10. Is there a discharge from the waterbody? D yes O no If yes. and the information is available, identify from the list
below the environment into which the waterbody discharges.
D River/Stream/Creek D onsite D offsite Distance
D Groundwater D onsite D offsite
D Wetland D onsite D offsite Distance
D Impoundment D onsite D offsite
11. Identify any field measurements and observations of water quality that were made. For those parameters for which
data were collected provide the measurement and the units of measure below:
Area
Depth (average)
Temperature (depth of the water at which the reading was taken)
pH
Dissolved oxygen
Salinity
Turbidity (clear, slightly turbid, turbid, opaque) (Secchi disk depth.
Other (specify)
12. Describe observed color and area of coloration.
13. Mark the open-water, non-flowing system on the site map attached to this checklist
40
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14. What observations, if any, were made at the waterbody regarding the presence and/or absence of benthic
rnacroinvertebrates, fish, birds, mammals, etc.?
41
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IV. AQUATIC HABITAT CHECKLIST - FLOWING SYSTEMS
Note: Aquatic systems are often associated with wetland habitats. Please refer to Section V. Wetland Habitat
Checklist.
1. What type(s) of flowing water system(s) is (are) present at the site?
D River D Stream D Creek
D Dry wash D Arroyo D Brook
D Artificially D Intermittent Stream D Channeling
created D Other (specify) ;
(ditch, etc.)
2. If known, what is the name of the waterbodv?
3. For natural systems, are there any indicators of physical alteration (e.g., channeling, debris, etc.)?
D yes D no If yes, please describe indicators that were observed.
4. What is the general composition of the substrate? Check all that apply.
D Bedrock D Sand (coarse) D Muck (fine/black)
D Boulder (>10 in.) D Silt (fine) D Debris
D Cobble (2.5-10 in.) D Marl (shells) D Detritus
C Gravel (0.1-2.5 in.) D Clay (slick) D Concrete
C Other (specify) .
5. What is the condition of the bank (e.g., height, slope, extent of vegetative cover)?
6. Is the system influenced by tides? O yes D no What information was used to make this determination?
42
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7. Is the flow intermittent? Dyes Dno If yes, please note the information that was used in making this determination.
8. Is there a discharge from the site to the waterbody? Dyes Dno If yes. please describe the discharge and its path.
9. Is there a discharge from the waterbody? D yes D no If yes. and the information is available, please identify what
the waterbody discharges to and whether the discharge is on site or off site.
10. Identify any field measurements and observations of water quality that were made. For those parameters for which
data were collected, provide the measurement and the units of measure in the appropriate space below:
Width (ft.)
Depth (ft.)
Velocity (specify units):
Temperature (depth of the water at which the reading was taken.
pH
Dissolved oxygen
Salinity ^ ,
Turbidity (clear, slightly turbid, turbid, opaque)
(Secchi disk depth ' )
Other (specify)
43
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11. Describe observed color and area of coloration.
12. Is any aquatic vegetation present? Dyes Ono If yes, please identify the type of vegetation present if known.
D Emergent D Submergent D Floating
13. Mark the flowing water system on the attached site map.
14. What observations were made at the waterbody regarding the presence and/or absence of benthic
macroinvenebrates, fish, birds, mammals, etc.?
44
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V. WETLAND HABITAT CHECKLIST
1. Based on observations and/or available information, are designated or known wetlands definitely present at the site?
Dyes Dno
Please note the sources of observations and information used (e.g., USGS Topographic Maps, National Wetland
Inventory. Federal or State Agency, etc.) to make this determination.
2. Based on the location of the site (e.g., along a waterbody, in a floodplain) and site conditions (e.g., standing water.
dark, wet soils; mud cracks; debris line; water marks), are wetland habitats suspected?
Dyes D no If yes, proceed with the remainder of the wetland habitat identification checklist
3. What type(s) of vegetation are present in the wetland?
D Submergent D Emergent
D Scrub/Shrub D Wooded
D Other (specify)
4. Provide a general description of the vegetation present in and around the wetland (height, color, etc.). Provide a
photograph of the known or suspected wetlands, if available.
5. Is standing water present? Dyes O no If yes, is this water O Fresh D Brackish
What is the approximate area of the water (sq. ft)? ;
Please complete questions 4,11,12 in Checklist ID • Aquatic Habitat - Non-Flowing Systems.
6 Is there evidence of flooding at the site? What observations were noted?
~ Buttressing D Water marks D Mud cracks
Z Debris line D Other (describe below)
45
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7. If known, what is the source of the water in the wetland?
D StreanVRiver/Creek/Lake/Pond D Groundwater
D Flooding D Surface Runoff
8. Is there a discharge from the site to a known or suspected wetland? D yes D no If yes, please describe.
. 9. Is there a discharge from the wetland? D yes D no. If yes, to what waterbody is discharge released?
G Surface Stream/River D Groundwater D Lake/Pond D Marine
10. If a soil sample was collected, describe the appearance of the soil in the wetland area. Circle or write in the best
response.
Color (blue/gray, brown, black, mottled) '
Water content (dry. wet. saturated/unsaturated).
11. Mark the observed wetland area(s) on the attached site map.
46
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APPENDIX B - Example of Flow Diagram For Conceptual Site Model
Figure B-1
Migration Routes of a Gas Contaminant
from Origin to Receptor
Original atal
of contaminant
of concern*
Pathway
from
origin
condenaatlon
**
Air
solidification
Change of
contaminant
atate In
pathway
n
Liquid —
_ **
f^*»*»
UaS
Solid
\j\jt\\j
n
Final
pathway
to receptor
,- > SO
1— > SW
1 — » so
•. AT
- • • > Ml
. QW
* OVY
a. QO
r oU
_> SW
r O VV
!
Human
_G'D
G,D
I,D
I,D
G,D
G,D
G,D
Receptor
Ecological Threat
Terrestrial
G,p
G,D
I,D
I,D
I,D
G,D
G,D
Aquatic
N/A
G,D
N/A
N/A
G,D
N/A
G.D
* M_
** Inc
e a transformation product
IBS vapors
Receptor Key
N/A
• Dermal Contact
- Inhalation
ittton
Applicable
• lnfl0$Uon
- Not Appld
Pathway Key
so - Son
SW - Surface Water
(Including aedlmenta)
QW • Ground Water
47
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Figure B-2
Migration Routes of a Liquid Contaminant
from Origin to Receptor
Original slate
of contaminant
of e neern*
Liquid
Pathway
from
origin
Change of
contaminant
state In
pathway
Final
pathway
to receptor
SW
crystallization
Liquid >
Gas" -[^
Solid »
SO
laachate,
Infiltration
Liquid
AI
* May be a transformation product
** Includes vapors
Gas
**
SW
AI
SW
SW
SO
SW
GW
> SO
•> AI
-> sw
Pathway Key
AI . AIT
30-Sod
SW • Surtae* Witw
(kiehidlng ledlmtnti)
QW - Qround Water
Receptor
Human
G,D
b*>
G^D
G,D
Ecological Threat
Terrestrial Aquatic
G,D G,p
N/A
J3,A
G,D
G,J)
G,D
G,D
oTo
G,D
G,D
G,D
N/A
N/A
Q,D
N/A
G,D
I,D
G,D
G,D
I,D
G,D
N/A
N/A
G,D
Recaptor K»y
D • Dwnui ConMel
I - tnhilitlon
Q Im---•!-.—
• mgvnrun
N/A - Not Appncabl*
48
-------
Figure B-3
Migration Routes of a Solid Contaminant
from Origin to Receptor
Original atate
of contaminant
of concern*
Solid
Pathway
from
origin
Change of
contaminant
•tale In
pathway
AI
partlculatea/
dual
Solid
SW
» Solid
> Liquid
**
so
Liquid
* May be a transformation product
** Includes vapors
jjeceptor Key
0 - (Mmd ContMl
I - Inhilrion
0 • (ng»«"on
N/A - Not ApplkaMi
Final
pathway
to receptor
AI
sw
so
_
Gas —
Solid
SO
AI
SW
SO
SO
GW
SW
Pathway Key
• Air
AI
SO
BW - Surfie* Wilw
flndudkiB Mdhntnti)
OW - Ground Wittr
Receptor
Human
1,0
Q,D
G,D
Ecological Threat
Terrestrial
1,0
G.D
G,D
Aquatic
N/A
6,0
N/A
Q,D
,0 G,D
G,D
G,D
G,D
1,0
G.D
G,D
G,D
G,D
G,D
G,D
1,0
G.D
G,D
G,D
N/A
G,D
N/A
N/A
G,D
N/A
N/A
N/A
G,D
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APPENDIX C - EXAMPLE SITES
Example sites are presented in this document to demonstrate how information from the checklist for ecological
assessment/sampling is used in conjunction with representative biological sampling to meet the study objectives. A
genera] history for each site is presented first, then additional preliminary information
I. SITE HISTORIES
Site A — Copper Site
This is a former municipal landfill located in an upland area of the mid-Atlantic plain. Residential, commercial, and
industrial refuse were disposed at the site from 1961 to 1980. Large amounts of copper wire were also disposed at this
site. Minimal grass cover has been placed over the fill. Terrestrial ecosystems in the vicinity of the landfill include
upland forest, successional fields, agricultural land, and residential and commercial areas. The surface of the landfill has
deteriorated in several locations. Leachate seeps have been noted on the slope of the landfill, several of which discharge
to a 5-acrc pond down-gradient of the site.
Site B - Stream DDT Site
This is a former chemical production facility located adjacent to a stream. The facility manufactured and packaged
dichlorodiphenyltrichloroethane (DDT). Due to poor storage practices, several DDT spills have occurred.
Site C — Terrestrial PCB Site
This site is a former waste oil recycling facility located in a remote area. Oils contaminated with polychlorinated
biphenyl compounds (PCBs) were disposed in a lagoon. The lagoon is not lined and the substrate is composed mostly of
sand. Oils contaminated with PCBs have migrated through the soil and contaminated a wide area adjacent to the site.
n. USE OF THE CHECKLIST FOR ECOLOGICAL ASSESSMENT/SAMPLING
Site A - Copper Site
A preliminary site visit was conducted, and the checklist indicated the following: 1) the pond has an organic substrate,
2) emergent vegetation including cattail and Phragmites occurs along the shore near the leachate seeps, and 3) the pond
reaches a depth of five feet toward the middle. Several species of sunfish. minnows, and carp were observed. A diverse
benthic macroinvertebrate community also has been noted in the pond. The pond appears to function as a valuable
habitat for fish and other wildlife.
Preliminary sampling indicated elevated copper levels in the seep as well as elevated base cations, total organic carbon
(TOO. and depressed pH levels (pH 5.7).
Copper can cause toxic effects in both aquatic plants and invertebrates at relatively low water concentrations, thereby
affecting the pond s ability to support macroinvertebrate and fish communities, as well as the wildlife that feed at the
pond Terrestrial ecosystems do not need to be evaluated because the overland flow of the seeps is limited to short
gullies. Thus, the area of concern has been identified as the 5-acre pond and the associated leachate seeps.
A review of the literature on the ecotoxicity of copper to aquatic biota and plants, both algae and vascular, was
conducted. In general it was found that young organisms are more sensitive to copper with decreasing sensitivity as
body weight increases. The toxiciry of copper in water is influenced by water hardness, alkalinity, and pH.
50
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Site B - Stream DDT Site
The ecological checklist was completed as part of the preliminary site visit The information gathered indicates that
surface water drainage from the site flows through several drainage swales toward a small unnamed creek. This creek is
a second order stream containing riffle-run areas and small pools. The stream substrate is composed of sand and gravel
in the pools with some small depositional areas in the backwater areas, and primarily cobble in the riffles. Previous
sampling efforts have indicated the presence of DDT and its metabolites in the stream sediments at a concentration of
230 milligrams per kilogram (mg/kg). A variety of wildlife, especially piscivorous birds, utilize this area for feeding.
Many species of minnow have been noted in this stream. DDT is well known for its tendency to bioaccumulate and
biomagnify in food chains, and available evidence indicates that it can cause reproductive failure in birds due to eggshell
thinning.
In freshwater systems, DDT can have direct effects on animals, particularly insects. A literature review of the aquatic
toxicity of DDT was conducted, and a no observed adverse effects level (NOAEL) was identified for aquatic insects.
Aquatic plants are not affected by DDT. Additional information on the effects of DDT on birds identified decreased
reproductive success due to eggshell thinning.
Site C — Terrestrial PCB Site
During a preliminary site visit, the ecological checklist was completed. Most of the habitat is upland forest, old field, -
and successions] terrestrial areas. Biological surveys at this site have noted a variety of small mammals, and red-tailed
hawks were also observed. The area of concern has been identified as the 10-acre area surrounding the site. PCBs have
been shown to reduce reproductive success in mammals or target liver functions. PCBs are not highly volatile, so
inhalation of PCBs would not be an important exposure pathway. However, PCBs have been shown to biomagnify
indicating that the ingestion exposure route needs evaluation. Shrews and/or voles would be appropriate mammalian
receptors to evaluate for this exposure route. Potential reproductive effects on predators that feed on small mammals
would also be important to evaluate. The literature has indicated that exposure to PCBs through the food chain can
cause chronic toxicity to predatory birds.
Limited information was available oh the effects of PCBs to red-tailed hawks. Studies on comparable species have
indicated decreased sperm concentration that may affect reproductive success.
ID. CONCEPTUAL SITE MODEL FORMULATION
Site A -- Copper Site
The assessment endpoint for this site was identified as the maintenance of pond fish and invertebrate community
composition similar to that of other ponds in the area of similar size and characteristics. Benthic macroinvertebrate
community studies may be relatively labor-intensive and potentially an insensitive measure in this type of system.
Measuring the fish community would also be unsuitable due to the limited size of the pond and the expected low
diversity of fish species. In addition, copper is not strongly food-chain transferrable. Therefore, direct toxicity testing
was selected as an appropriate measurement endpoint. Toxicity was defined as a statistically significant decrease in
survival or juvenile growth rates in a population exposed to water or sediments, as compared to a population from the
reference sites.
One toxicity test selected was a 10-day solid-phase sediment toxicity test using early life-stage Hyalella azteca. The
measurement endpomis for the test are mortality and growth rates (measured as length and weight changes). Two water-
column toxicity tests were selected: a 7-day test using the alga Selenastrum capricornutum (growth test) and a 7-day
larval fish test using Pimephales promelas (mortality and growth endpoints).
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Five sediment samples were collected from the pond bottom at intervals along an identified concentration gradient
Reference sediment was also collected. A laboratory control was utilized in addition to the reference sediment in this
toxiciry test The study design specified that sediment for the toxicity tests was collected from the leachate seeps known
to be at the pond edge, and from four additional locations transecting the pond at equidistance locations. A pre-sampling
visit was required to confirm that the seep was flowing due to the intermittent nature of leachate seeps.
Site B — Stream DDT Site
A conceptual model was developed to evaluate the environmental pathways for DDT that could result in ecological
impacts. DDT in the sediments can be released to the water column during natural resuspension and redistribution of
the sediments. Some diffusion of DDT to the water column from the sediment surface may also occur. The benthic
macroinvertebrate community would be an initial receptor for Vie DDT in sediments. Fish that feed on the benthic
macroin vertebrates could be exposed to the DDT both in the water column and in their food. Piscivorous birds would
be exposed to the DDT that has accumulated in the fish. For example, belted kingfishers are known to feed in the
stream. Given the natural history of this species, it is possible that they forage entirely in the contaminated area. From
this information, the assessment endpoint was identified to be the protection of piscivorous birds from eggshell thinning
due to DDT exposure. From this assessment endpoint eggshell thinning in the belted kingfisher was selected as the
measurement endpoint. .
Existing information identified a DDT gradient in the stream sediments. Forage fish (e.g., creek chub) were selected to ,
measure exposure levels for kingfishers. The study design for measuring DDT residue levels specified that 10 creek
chub of the same size and sex will be collected at each location for chemical residue analysis. Although analytical data
for the stream sediment exists, new co-located sediment samples were specified to be collected to provide a stronger
link between the present state of contamination in the sediment and in the fish.
Site C - Terrestrial PCB Site
A conceptual model was prepared to determine the exposure pathways by which predatory birds could be exposed to
PCBs originating in the soil at the site. The prey of red-tailed hawks includes voles, deer mice, and various insects.
Voles are herbivorous and prevalent at the site. However. PCBs do not strongly accumulate in plants, thus voles may
not represent a strong exposure pathway to hawks. Deer mice are omnivorous and may be more likely than voles to be
exposed to PCBs. The assessment endpoint for this site was identified to be the protection of reproductive success in
high trophic level species exposed to PCBs via diet
Initially, a sampling feasibility study was conducted to confirm sufficient numbers of the deer mice. Two survey lines of
10 live traps were set for deer mice in the area believed to contain the desired concentration gradient for the study
design. Previous information indicated a gradient of decreasing PCB concentration with increasing distance from the
unlined lagoon. Three locations were selected along this gradient to measure PCB concentrations in prey. Co-located
soil and water samples were also collected. The analytical results of these matrices were utilized as variables in a food
chain accumulation model which predicted the amount of contaminant in the environment that may travel through the
food chain, ultimately to the red-tailed hawk. . .
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REFERENCES
ASTM. 1992. Standard Guide for Conducting Early Life-Stage Toxicity Tests with Fishes. American Society for
Testing and Materials. £1241-92.
Bligh. E.G., WJ. Dyer. 1959. Ltpid Extraction and Purification. Canadian Journal of Biochemistry and Physiology. Vol
37. pp. 912-917
Brungs, W.A. and D.I. Mount 1978. Introduction to a Discussion of the Use of Aquatic Toxiciry Tests for Evaluation
of the Effects of Toxic Substances. Cairns, J. Jr., ILL. Dickson and A.W. Makei (eds.) Estimating the Hazard of
Chemical Substances to Aquatic Life. ASTM 657. Amer. Soc. Test Materials, Philadelphia, PA. p. 1526.
Green, J.C., CL. Baucis, WJ. Warren-Hicks, B.R. Parkhurst, GJ-. Under, S.A. Peterson, and W£. Meiller. 1989.
Protocol for Short Term Toxiciry Screening of Hazardous Waste. VS. Environmental Protection Agency,
Environmental Research Laboratory, Corvallis, OR. EPA 600/3-88/029.
Hair, JD. 1980. Measurement of ecological diversity, in S.D. Schemnitz, ed. Wildlife Management Techniques
Manual. Fourth Edition. The Wildlife Society, Washington, D.C. pp269-275.
Hayes, MX. 1983. Active Fish Capture Methods, Chapter 7 in Fisheries Techniques. American Fisheries Society, pp.
123-145. .
Heroes, S.E. and C.P. Allen. 1983. Lipid Quantification of Freshwater Invertebrates: Method Modification for
Microquanritation. Canadian Journal of Fisheries and Aquatic Sciences. 40(8). pp. 1315-1317.
Hurben, W.A. 1983. Passive Capture Methods, Chapter 6 in Fisheries Techniques. American Fisheries Society, pp. 95-
122.05
Karr, J.R., K.D. Fausch, PI.. Angermeier. P.R. Yam, and U. Schlosser. 1986. Assessing Biological Integrity in
Running Waters: A Method and Its Rationale. Special Publication 5. Illinois Natural History Survey.
Philips. DJ.H. 1977. The Use of Biological Indicator Organisms to Monitor Trace Metal Pollution In Marine and
Estuarine Environments-A Review. Environmental Poll 13, pp. 281-317.
Philips. DJ.H. 1978. Use of Biological Indicator Organisms to Quantitate Organochlorine Pollutants in Aquatic
Environments-A Review. Environmental Poll. 16, pp. 167-229.
Timbrell. J.A. 1989. Introduction to Toxicology. Taylor and Francis, London. 155p.
U.S. EPA (Environmental Protection Agency). 1997. Ecological Risk Assessment Guidance for Superfund: Process for
Designing and Conducting Ecological Risk Assessments. Office of Solid Waste and Emergency Response. EPA 540-R-
97/006.
U.S. EPA (Environmental Protection Agency). 1994. CLP National Functional Guidelines for Inorganic Data .
Review Office of Solid Waste and Emergency Response. Publication 9240.1-05
U.S. EPA (Environmental Protection Agency). January 1991. Compendium ofERTToxicity Testing Procedures.
OSWER Directive 9360.4-08.
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U.S. EPA (Environmental Protection Agency). 1992. Framework for Ecological Risk Assessment. EPA/630/R-92/001.
U.S. EPA (Environmental Protection Agency). December 1991b. ECO Update. Volume 1, Number 2, Publication
9345.0-051. Office of Emergency and Remedial Response, Hazardous Site Evaluation Division (OS-230).
U.S. EPA (Environmental Protection Agency). .April 1990. Quality Assurance/Quality Control (QA/QC) Guidance
for Removal Activities. Sampling QA/QC Plan and Data Validation Procedures. EPA/540/G-90/004.
U.S. EPA (Environmental Protection Agency). November 1990. Macroinvertebrate Field and Laboratory Methods
for Evaluating the Biological Integrity of Surface Waters.. Aquatic Biology Branch and Development and Evaluation
Branch, Quality Assurance Research Division, Environmental Monitoring Systems Laboratory, Cincinnati, Ohio,
EPA/600/4-90/030.
U.S. EPA (Environmental Protection Agency). March 1989b. Short-Term Methods for Estimating the Chronic Toxicity
of Effluents and Receiving Waters to Freshwater Organisms. EPA/600/4-89/001.
U.S. Environmental Protection Agency. May I989a. Rapid Bioassessment Protocols For Use In Streams And Rivers:
Benthic Macroinvertebrates and Fish. EPA/444/4-89-001.
U.S. Environmental Protection Agency. May 1988. Short-Term Methods for Estimating the Chronic Toxicity of
Effluents and Receiving Waters to Marine and Estuarine Organisms. EPA/600/4-87/028.
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APPENDIX C
SUPPLEMENTAL GUIDANCE ON LITERATURE SEARCH
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APPENDIX C
SUPPLEMENTAL GUIDANCE ON LITERATURE SEARCH
A literature search is conducted to obtain information on contaminants of concern,
their potential ecological effects, and species of concern. This appendix is separated into two
sections; Section C-l describes the information necessary for the literature review portion of
an ecological risk assessment. Topics include information for exposure profiles,
bioavailability or bioconcentration factors for various compounds, life-history information for
the species of concern or the surrogate species, and an ecological effects profile. Section C-2
lists information sources and techniques for a literature search and review. Topics include a
discussion of how to select key words on which to base a search and various sources of
information (i.e., databases, scientific abstracts, literature reviews, journal articles, and
government documents). Threatened and endangered species are discussed separately due to
the unique databases and information sources available for these species.
Prior to conducting a literature search, it is important to determine what information is
needed for the ecological risk assessment. The questions raised in Section D-l must be
thoroughly reviewed, the information necessary to complete the assessment must be
determined, and the purpose of the assessment must be clearly defined. Once these activities
are completed, the actual literature search can begin. These activities will assist in focusing
and streamlining the search.
C-1 LITERATURE REVIEW FOR AN ECOLOGICAL RISK ASSESSMENT
Specific information. During problem formulation, the risk assessor must
determine what information is needed for the risk assessment. For example, if the risk
assessment will estimate the effects of lead contamination of soils on terrestrial vertebrates,
then literature information on the effects of dissolved lead to fish would not be relevant. The
type and form of the contaminant and the biological species of concern often can focus the
literature search. For example, the toxicity of organometallic compounds is quite different
from the comparable inorganic forms. Different isomers of organic compounds also can have
different toxic effects.
Reports of toxicity tests should be reviewed critically to ensure that the study was
scientifically sound. For example, a report should specify the exposure routes, measures of
effect and exposure, and the full study design. Moreover, whether the investigator used
accepted scientific techniques should be determined.
The exposure route used in the study should also be comparable to the exposure route
in the risk assessment. Data reported for studies where exposure is by injection or gavage are
not directly comparable to dietary exposure studies. Therefore, an uncertainty factor might
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need to be- included in the risk assessment study design, or the toxicity report should not be
used in the risk assessment.
To use some data reported in the literature, dose conversions are necessary to estimate
toxicity levels for species other than those tested. Doses for many laboratory studies are
reported in terms of mg contaminant/kg diet, sometimes on a wet-weight basis and sometimes
on a dry-weight basis. That expression should be converted to mg contaminant/kg wet
bodyweight/day, so that estimates of an equivalent dose in another species can be scaled
appropriately. Average ingestion rate and wet body weight for a species often are reported in
the original toxicity study. If not, estimates of those data can be obtained from other
literature sources to make the dose conversion:
Dose = (mg contaminant/kg diet) x ingestion rate (kg/day) x (I/wet body weight (kg)).
If the contaminant concentration is expressed as mg contaminant/kg dry diet, the ingestion
rate should also be in terms of kg of dry diet ingested per day.
Exposure profile. Once contaminants of concern are selected for the ecological risk
assessment, a general overview of the contaminants' physical and chemical properties is
needed. The fate and transport of contaminants in the environment determines how biota are
likely to be exposed. Many contaminants undergo degradation (e.g., hydrolysis, photolysis,
microbial) after release into the environment. Degradation can affect toxicity, persistence,
and fate and transport of compounds. Developing an exposure profile for a contaminant
requires information regarding inherent properties of the contaminant that can affect fate and
transport Or bioavailability.
Bioavailability. Of particular importance in an ecological risk assessment is the
bioavailability of site contaminants in the environment. Bioavailability influences exposure
levels for the biota. Some factors that affect bioavailability of contaminants in soil and
sediment include the proportion of the medium composed of organic matter, grain size of the
medium, and its pH. The aerobic state of sediments is important because it often affects the
chemical form of contaminants. Those physical properties of the media can change the
chemical form of a contaminant to a form that is more or less toxic than the original
contaminant. Many contaminants adsorb to organic matter, which can make them less
bioavailable. •
Environmental factors that influence the bioavailability of a contaminant in water are
important to aquatic risk assessments. Factors including pH, hardness, or aerobic status can
determine both the chemical form and uptake of contaminants by biota. Other environmental
factors can influence how organisms process contaminants. For example, as water
temperatures rise, metabolism of fish and aquatic invertebrates increases, and the rate of
uptake of a contaminant from water can increase.
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If the literature search on the contaminants of concern reveals information on the
bioavailability of a contaminant, then appropriate bioaccumulation or bioconcentration factors
(BAFs or BCFs) for the contaminants should be determined. If not readily available in the
literature, BAF or BCF values can be estimated from studies that report contaminant
concentrations in both the environmental exposure medium (e.g., sediments) and in the
exposed biota (e.g., benthic macroinvertebrates). Caution is necessary, however, when
extrapolating BAF or BCF values estimated for one ecosystem to another ecosystem.
Life history. Because it is impossible and unnecessary to model an entire ecosystem,
the selection of assessment endpoints and associated species of concern, and measurement
endpoints (including those for a surrogate species if necessary) are fundamental to a
successful risk assessment. This process is described in Steps 3 and 4. Once assessment and
measurement endpoints are agreed to by the risk assessor and risk manager, life history
information for the species of concern or the surrogate species should be collected. Patterns
of activity and feeding habits of a species affect their potential for exposure to a contaminant
(e.g., grooming activities of small mammals, egestion of bone and hide by owls). Other
important exposure factors include food and water ingestion rates, composition of the diet,
average body weight, home range size, and seasonal activities such as migration.
Ecological effects profile. Once contaminants and species of concern are selected
during problem formulation, a general overview of toxicity and toxic mechanisms is needed.
The distinction between the species of concern representing an assessment endpoint and a
surrogate species representing a measurement endpoint is important. The species of concern
is the species that might be threatened by contaminants at the site. A surrogate species is
used when it is not appropriate or possible to measure attributes of the species of concern. A
surrogate for a species of concern should be sufficiently similar biologically to allow
inferences on likely effects in the species of concern.
The ecological effects profile should include toxicity information from the literature
for each possible exposure route. A lowest-observed-adverse-effect level (LOAEL) and the
no-observed-adverse-effect level (NOAEL) for the species of concern or its surrogate should
be obtained. Unfortunately, LOAELs are available for few wildlife species and contaminants.
If used with caution, toxicity data from a closely related species can be used to estimate a
LOAEL and a NOAEL for a receptor species.
C-2 INFORMATION SOURCES
This section describes information sources that can be examined to find the
information described in Section 3-1. A logical and focused literature search will reduce the
time, spent searching for pertinent information.
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A first step in a literature search is to develop a search strategy, including a list of key
words. The next step is to review computerized databases, either on-line or CD-ROM-based
information systems. These systems can be searched based on a number of parameters.
Scientific abstracts that contain up-to-date listings of current, published information
also are useful information sources. Most abstracts are indexed by author or subject.
Toxicity studies and information on wildlife life-histories often are summarized in literature
reviews published in books or peer-reviewed journals. Original reports of toxiciry studies can
be identified in the literature section of published documents. The original article in which
data are reported must be reviewed before the data are cited in a risk assessment.
Key words. Once the risk assessor has prepared a list of the specific information
needed for the risk assessment, a list of key words can be developed. Card catalogs,
abstracts, on-line databases, and other reference materials usually are indexed on a limited set
of key words. Therefore, the key words used to search for information must be considered
carefully.
Useful key words include the contaminant of concern, the biological species of
concern, the type of toxicity information wanted, or other associated words. In addition,
related subjects can be used as key words. However, it usually is necessary to limit
peripheral aspects of the subject in order to narrow the search. For example, if the risk
assessor needs information on the toxicity of lead in soils to moles, then requiring that both
"lead" and "mole" are among the key words can focus the literature search. If the risk
assessor needs information on a given plant or animal species (or group of species), key
words should include both the scientific name (e.g., genus and species names or order or
family names) and an accepted common name(s). The projected use of the data in the risk
assessment helps determine which key words are most appropriate.
If someone outside of the risk assessment team will conduct the literature search, it is
important that they understand both the key words and the study objectives for the data.
Databases. Databases are usually on-line or CD-ROM-based information systems.
These systems can be searched using a number of parameters. Prior to searching databases,
the risk assessor should determine which database(s) is most likely to provide the information
needed for the risk assessment. For example, U.S. Environmental Protection Agency's
(EPA's) AQUIRE database (AQUatic Information REtrieval database) provides information
specifically on the toxicity of chemicals to aquatic plants and animals. PHYTOTOX includes
data on the toxicity of contaminants to terrestrial and aquatic plants, and TERRETOX
includes data on toxicity to terrestrial animals. U.S. EPA's IRIS (Integrated Risk Information
System) provides information on human health risks (e.g., references to original toxicity
studies) and regulatory information (e.g., reference doses and cancer potency factors) for a
variety of chemicals. Other useful databases include the National Library of Medicine's
HSDB (Hazardous Substances Data Bank) and the National Center for Environmental
Assessment's HEAST Tables (Health Effects Assessment Summary Tables). Commercially
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available databases include BIOSIS (Biosciences Information Services) and ENVIROLINE.
Another database, the U.S. Public Health Service's Registry of Toxic Effects of Chemical
Substances (RTECS) is a compilation of toxicity data extracted from the scientific literature
and is also available online.
Several states have Fish and Wildlife History Databases or Academy of Science
databases, which often provide useful information on the life-histories of plants and animals
in the state. State databases are particularly useful for obtaining information on endemic
organisms or geographically distinct habitats.
Databases searches can yield a large amount of information in a short period of time.
Thus, if the key words do not accurately describe the information needed, database searches
can provide a large amount of irrelevant information. Access fees and on-line fees can apply;
therefore, the selection of relevant key words and an organized approach to the search will
reduce the time and expense of on-line literature searches.
Abstracts. Published abstract compilations (e.g., Biological Abstracts, Chemical
Abstracts, Applied Ecology Abstracts) contain up-to-date listings of current, published
information. Most abstracts are indexed by author or subject. Authors and key words can be
cross-referenced to identify additional publications. Abstract compilations also include, for
each citation, a copy of its abstract from the journal or book in which it was published.
Reviewing the abstracts of individual citations is a relatively quick way to determine whether
an article is applicable to the risk assessment. As with computerized database searches, it is
important to determine which abstract compilations are most suitable for the risk assessor's
information needs.
Published abstract compilations that are indexed by author are particularly useful. If
an author is known to conduct a specific type of research, their name would be referenced in
the abstract for other articles on similar subjects. If the risk assessor considers an abstract
pertinent to the assessment, the original article must be retrieved and reviewed before it can
be cited in the risk assessment. Otherwise, the results of the risk assessment could be based
on incorrect and incomplete information about a study.
Abstracts usually must be searched manually, which can be a very time consuming.
The judicious use of key words can help to reduce the amount of time needed to search
through these volumes.
Literature review publications. Published literature reviews often cover toxicity
or wildlife information of value to an ecological risk assessment. For example, the U.S. Fish
and Wildlife Services (U.S. FWS) has published several contaminant-specific documents that
list lexicological data on terrestrial, aquatic, and avian studies (e.g., Eisler, 1988). The U.S.
EPA publishes ambient water quality criteria documents (e.g., U.S. EPA, 1985) that list all
the data used to calculate those values. Some literature reviews critically evaluate the original
studies (e.g., toxicity data reviewed by NOAA, 1990). The Wildlife Exposure Factors
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Handbook (U.S. EPA, 1993a,b) provides pertinent information on exposure factors (e.g., body
weights, food ingestion rates, dietary composition, home range size) for 34 selected wildlife
species.
Literature reviews can provide an extensive amount of information. However, the risk
assessor must obtain a copy of the original of any studies identified in a literature review that
will be used in the risk assessment. The original study must be reviewed and evaluated
before it can be used in the risk assessment. Otherwise, the results of the risk assessment
could be based on incorrect and incomplete information about a study.
References cited in previous studies. Pertinent studies can be identified in the
literature cited section of published documents that are relevant to the risk assessment, and
one often can identify several investigators who work on related studies. Searching for
references in the literature cited section of published documents, however, takes time and
might hot be very effective. However, this is probably the most common approach to
identifying relevant literature. If this approach is selected, the best place to start is a review
article. Many journals do not list the title of a citation for an article, however, limiting the
usefulness of this technique. Also, it can be difficult to retrieve literature cited in obscure or
foreign journals or in unpublished masters' theses or doctoral dissertations. Although this
approach tends to be more time consuming than the other literature search approaches
described above, it probably is the most common approach used to locate information for a
risk assessment.
Journal articles, books, government documents. There are a variety of
journals, books, and government documents that contain information useful to risk
assessments. The same requirement for retrieving the original reports for any information
used in the risk assessment described for other information sources applies to these sources.
Threatened and endangered species. Threatened and endangered species are of
concern to both federal and state governments. When conducting an ecological risk
assessment, it often is necessary to determine or estimate the effects of site contaminants to
federal threatened or endangered species. In addition, other special-status species (e.g.,
species listed by a state as endangered or threatened within the state) also can be the focus of
the assessment. During the problem formulation step, the U.S. FWS or state Natural Heritage
programs should be contacted to determine if these species are present or might be present on
or near a Superfund site.
Once the presence of a special-status species is confirmed or considered likely,
information on this species, as well as on surrogate species, should be included in the
literature search. There are specific federal and state programs that deal with issues related to
special-status species, and often there is more information available for these than for non-
special-status species used as surrogates for an ecological risk assessment. Nonetheless, the
use of surrogate species usually is necessary when an assessment endpoint is a special-status
species.
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REFERENCES
Eisler, R. 1988. Lead Hazards to Fish, Wildlife, and Invertebrates: A Synoptic Review.
U.S. Fish and Wildlife Service Patuxent Wildlife Research Center, Laurel MD: U.S.
Department of the Interior; Biological Report 85(1.14), Contaminant Hazard Reviews
Rep. No. 14.
National Oceanic and Atmospheric Administration (NOAA). 1990. The Potential for
Biological Effects of Sediment-Sorbed Contaminants Tested in the National Status and
Trends Program. Seattle, WA: Office of Oceanography and Marine Assessment.
NOAA/TM/NOS/6MA-52. Technical memorandum by Long, E.R. and Morgan, L.G.
U.S. Environmental Protection Agency (U.S. EPA). 1993a. Wildlife Exposure Factors
Handbook Volume I. Washington, DC: Office of Research and Development;
EPA/600/R-93/187a.
U.S. Environmental Protection Agency (U.S. EPA). 1993b. Wildlife Exposure Factors
Handbook Volume II: Appendix. Washington, DC: Office of Research and
Development; EPA/600/R-93/187b.
U.S. Environmental Protection Agency (U.S. EPA). 1985. Ambient Water Quality Criteria
for Copper-1984. Washington, DC: Office of Water, Regulations and Standards,
Criteria and Standards Division. EPA/440/5-84-031. PB85-227023.
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APPENDIX D
STATISTICAL CONSIDERATIONS
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APPENDIX D
STATISTICAL CONSIDERATIONS
In the biological sciences, statistical tests often are needed to support decisions based
on alternative hypotheses because of the natural variability in the systems under investigation.
A statistical test examines a set of sample data, and, based on an expected distribution of the
data, leads to a decision on whether to accept the hypothesis underlying the expected
distribution or whether to reject that hypothesis and accept an alternative one. The null
hypothesis is a hypothesis of no differences. It usually is formulated for the express purpose
of being rejected. The alternative or test hypothesis is an operational statement of the
investigator's research hypothesis. An example of a null hypothesis for toxicity testing would
be that mortality of water fleas exposed to water from a contaminated area is no different
than mortality of water fleas exposed to water from an otherwise similar, but uncontaminated
area. An example of the test hypothesis is that mortality of water fleas exposed to water
from the contaminated area is higher than mortality of water fleas exposed to uncontaminated
water.
D-1 TYPE I AND TYPE II ERROR
There are two types of correct decisions for hypothesis testing: (1) accepting a true
null hypothesis, and (2) rejecting a false null hypothesis. There also are two types of
incorrect decisions: rejecting a true null hypothesis, called Type I error; and accepting a false
null hypothesis, called Type II error.
When designing a test of a hypothesis, one should decide what magnitude of Type I
error (rejection of a true null hypothesis) is acceptable. Even when sampling from a
population of known parameters, there are always some sample sets which, by chance, differ
markedly. If one allows 5 percent of samples to lead to a Type I error, then one would on
average reject a true null hypothesis for 5 out of every 100 samples taken. In other words,
we would be confident that, 95 times out of 100, one would not reject the null hypothesis of
no difference "by mistake" (because chance alone produced such deviant results). When the
probability of Type I error (commonly symbolized by a) is set at 0.05, this is called a
significance level of 5 percent. Setting a significance level of 5 percent is a widely accepted
convention in most experimental sciences, but it is just that, a convention. One can demand
more confidence (e.g., a = 0.01) or less confidence (e.g., a = 0.10) that the hypothesis of no
difference is not rejected by mistake.
If one requires more confidence for a given sample size that the null hypothesis is not
rejected by mistake (e.g., a = 0.01), the chances of Type II error increase. In. other words,
the chance increases that one will mistakenly accept a false null hypothesis (e.g., mistakenly
D-1
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believe that the contaminated water from the site has no effect on mortality of water fleas).
The probability of Type II error is commonly denoted by p. Thus:
p (Type I error) = a -
p (Typell error) = p
v
However, if one tries to evaluate the probability of Type n error (accepting a false hypothesis
of no difference), there is a problem. If the null hypothesis is false, then some other
hypothesis must be true, but unless one can specify a second hypothesis, one can't determine
the probability of Type n error. This leads to another important statistical consideration,
which is the power of a study design and the statistical test used to evaluate the results.
D-2 STATISTICAL POWER
The power of a statistical test is equal to (1 - P) and is equal to the probability of
rejecting the null hypothesis (no difference) when it should be rejected (i.e., it is false) and
the specified alternative hypothesis is true. Obviously, for any given test (e.g., a toxicity test
at a Superfund site), one would like the quantity (1 - P) to be as large as possible (and P to
be as small as possible). Because one generally cannot specify a given alternative hypothesis
(e.g., mortality should be 40 percent in the exposed population), the power of a test is
generally evaluated on the basis of a continuum of possible alternative hypotheses.
•
Ideally, one would specify both a and P before an experiment or test of the hypothesis
is conducted. In practice, it is usual to specify a (e.g., 0.05) and the sample size because the
exact alternative hypothesis cannot be specified.1 Given the inverse relationship between the
likelihood of making Type I and Type II errors, a decrease in a will increase P for any given
sample size.
To improve the statistical power of a test (i.e., reduce p), while keeping a constant,
one can either increase the sample size (N) or change the nature of the statistical test. Some
statistical tests are more powerful than others, but it is important that the assumptions
required by the test (e.g., normality of the underlying distribution) are met for the test results
to be valid. In general, the more powerful tests rely on more assumptions about the data (see
Section D-3).
Alternative study designs sometimes can improve statistical power (e.g., stratified
random sampling compared with random sampling if something is known about the history
and location of contaminant release). A discussion of different statistical sampling designs is
beyond the scope of this guidance, however. Several references provide guidance on
statistical sampling design, sampling techniques, and statistical analyses appropriate for
hazardous waste sites (e.g., see Cochran, 1977; Green, 1979; Gilbert, 1987; Ott, 1995).
' With a specified alternative hypothesis, once a and the sample size (N) are set, P is determined.
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One also can improve the power of a statistical test if the test hypothesis is more
specific than "two populations are different," and, instead, predicts the direction of a
difference (e.g., mortality in the exposed group is .higher than mortality in the control group).
When one can predict the direction of a difference between groups, one uses a one-tailed
statistical test; otherwise, one must use the less powerful two-tailed version of the test.
Highlight D-2
Key Points About Statistical Significance, Power, and Sample Size
(1) The significance level for a statistical test, a, is the probability that a statistical test will
'yield a value under which the null hypothesis will be rejected when it is in fact true.
In other words, a defines the probability of committing Type I error (e.g., concluding
that the site medium is toxic when it is in fact not toxic to the test organisms).
(2) The value of p is the probability that a statistical test will yield a value under which the
null hypothesis is accepted when it is in fact false. Thus, P defines the probability of
committing Type n error (e.g., concluding that the site medium is not toxic when it is
in fact toxic to the test organisms).
(3) The power of a statistical test (i.e., 1 - f3) indicates the probability of rejecting the null
hypotheses when it is false (and therefore should be rejected). Thus, one wants the
power of a statistical test to be as high as possible.
(4) Power is related to the nature of the statistical test chosen. A one-tailed test is more
powerful than a two-tailed test. If the alternative to the null hypothesis can state the
expected direction of a difference between a test and control group, one can use the more
powerrulone-tailed test.
(5) The power of any statistical test increases with increasing sample size.
D-3 STATISTICAL MODEL
Associated with every statistical test is a model and a measurement requirement. Each
statistical test is valid only under certain conditions. Sometimes, it is possible to test whether
the conditions of a particular statistical model are met, but more often, one has to assume that
they are or are not met based on an understanding of the underlying population and sampling
design. The conditions that must be met for a statistical test to be valid often are referred to
as the assumptions of the test.
The most powerful statistical tests (see previous section) are those with the most
extensive assumptions. In general, parametric statistical tests (e.g., t test, F test) are the most
powerful tests, but also have the most exacting assumptions to be met:
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(1) The "observations" must be independent;
(2) The "observations" must be drawn from a population that is normally
distributed;
(3) The populations must have the same variance (or in special cases, a known
ratio of variances); and
(4) The variables must have been measured at least on an interval scale so that it is
possible to use arithmetic operations (e.g., addition, multiplication) on the
measured values (Siegel, 1956).
The second and third assumptions are the ones most often violated by the types of data
associated with biological hypothesis testing. Often, distributions are positively skewed (i.e.,
longer upper than lower tail of the distribution). Sometimes, it is possible to transform data
from positively skewed distributions to normal distributions using a mathematical function.
For example, many biological parameters turn out to be log-normally distributed (i.e., if one
takes the log of all measures, the resulting values are normally distributed). Sometimes,
however, the underlying shape of the distribution cannot be normalized (e.g., it is bimodal).
When the assumptions required for parametric tests are not met, one must use
nonparametric statistics (e.g., median test, chi-squared test). Nonparametric tests are in
general less powerful than parametric tests because less is known or assumed about the shape
of the underlying distributions. However, the loss in power can be compensated for by an
increase in sample size, which is the concept behind measures of power-efficiency.
Power-efficiency reflects the increase in sample size necessary to make test B (e.g., a
nonparametric test) as efficient or powerful as test A (e.g., a parametric test). A power-
efficiency of 80 percent means that in order for test B to be as powerful as test A, one must
make 10 observations for test B for every 8 observations for test A.
For further information .on statistical tests, consult references on the topic (e.g., see
references below).
REFERENCES
Cochran, W. G. 1977. Sampling Techniques. Third edition. New York, NY: John Wiley
and Sons, Inc.
^
Gilbert, R.O. 1987. Statistical Methods for Environmental Pollution Monitoring. New York,
NY: Reinhold.
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Green, R. H. 1979. Sampling Design and Statistical Methods for Environmental Biologists.
New York, NY: Wiley.
Ott, W.R. 1995. Environmental Statistics arid Data Analysis. Boca Raton, FL: CRC Press,
Inc., Lewis Publishers.
Siegel, S. 1956. Non-parametric Statistics. New York, NY: McGraw-Hill.
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