United States
Environmental Protection
Agency
Office of Water
(4305)
EPA 823-B-00-008
November 2000
SEPA Guidance for Assessing
Chemical Contaminant Data
for Use in Fish Advisories
Volume 2
Risk Assessment and Fish
Consumption Limits
Third Edition
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Guidance for Assessing Chemical Contaminant
Data for Use in Fish Advisories
Volume 2: Risk Assessment and Fish Consumption Limits
Third Edition
Office of Science and Technology
Office of Water
U.S. Environmental Protection Agency
Washington, DC
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vvEPA
United States
Environmental Protection Agency
(4305)
Washington, DC 20460
Official Business
Penalty for Private Use $300
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c/EPA
Guidance for Assessing Chemical Contaminant Data for Use In Fish Advisories
Volume 2: Risk Assessment and Fish Consumption Limits
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TABLE OF CONTENTS
TABLE OF CONTENTS
Section Page
List of Figures vii
List of Tables viii
List of Acronyms x
Acknowledgments xiii
Executive Summary xiv
1 Introduction 1-1
1.1 Overview 1-2
1.2 Objectives 1-3
1.3 Sensitive Subpopulations 1-6
1.4 Contents of Volume 2 1-10
1.5 Changes to Volume 2 1-13
1.6 Sources 1-14
2 Risk Assessment Methods 2-1
2.1 Introduction 2-1
2.1.1 Other Information Sources 2-3
2.2 Hazard Identification 2-4
2.2.1 Approach for Fish Contaminants 2-5
2.2.2 Assumptions and Uncertainty Analysis 2-8
2.3 Dose-Response Assessment 2-10
2.3.1 Carcinogenic Effects 2-12
2.3.2 Noncarcinogenic Effects 2-13
2.3.3 Mutagenicity/Genotoxicity 2-19
2.3.4 Multiple Chemical Exposures: Interactive Effects 2-20
2.3.5 Assumptions and Uncertainties 2-22
2.4 Exposure Assessment 2-25
2.4.1 Chemical Occurrences in Fish 2-25
2.4.2 Geographic Distribution of Contaminated Fish 2-27
2.4.3 Individual Exposure Assessment 2-27
2.4.4 Population Exposure Assessments 2-33
2.4.5 Uncertainty and Assumptions 2-39
2.5 Risk Characterization 2-49
2.5.1 Carcinogenic Toxicity 2-51
2.5.2 Noncarcinogenic Toxicity 2-52
2.5.3 Subpopulation Considerations 2-53
2.5.4 Multiple Species and Multiple Contaminant
Considerations 2-55
in
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TABLE OF CONTENTS
Section Page
2.5.5 Incorporating Considerations of Uncertainty in
Consumption Limits 2-55
2.6 Summarizing Risk Data 2-55
Development and Use of Risk-Based Consumption Limits 3-1
3.1 Overview and Section Organization 3-1
3.2 Equations Used to Develop Risk-Based
Consumption Limits 3-2
3.2.1 Calculation of Consumption Limits for
Carcinogenic Effects 3-2
3.2.2 Calculation of Consumption Limits for
Noncarcinogenic Effects 3-7
3.2.3 Developmental Effects 3-8
3.3 Default and Alternative Values for Calculating Consumption
Limits 3-9
3.3.1 Maximum Acceptable Risk Level 3-10
3.3.2 Cancer Potencies and Chronic Reference Doses
(q/s and RfDs) 3-11
3.3.3 Consumer Body Weight (BW) 3-11
3.3.4 Meal Size 3-12
3.3.5 Contaminant Concentration in Fish Tissue 3-14
3.3.6 Modifying Time-Averaging Period (Tap) 3-14
3.4 Modification of Consumption Limits for a
Single Contaminant in a Multispecies Diet 3-15
3.4.1 Carcinogenic Effects 3-15
3.4.2 Noncarcinogenic Effects 3-16
3.5 Modification of Consumption Limits for
Multiple Contaminant Exposures 3-17
3.5.1 Carcinogenic Effects 3-19
3.5.2 Noncarcinogenic Effects 3-20
3.5.3 Species-Specific Consumption Limits in a Multiple
Species Diet 3-22
Risk-Based Consumption Limit Tables 4-1
4.1 Overview and Section Organization 4-1
4.2 Consumption Limit Tables 4-2
Toxicological Profile Summaries for Target Analytes 5-1
5.1 Introduction 5-1
5.1.1 Categories of Information Provided for Target Analytes .... 5-1
5.1.2 Abbreviations Used and Scientific Notation 5-8
5.2 Metals 5-9
5.2.1 Arsenic 5-9
5.2.2 Cadmium 5-13
5.2.3 Mercury 5-18
IV
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TABLE OF CONTENTS
Section
5.2.4 Selenium
5.2.5 Tributyltin Oxide
5.3 Organochlorine Pesticides
5.3.1 Chlordane
5.3.2 DDT, DDE, ODD
5.3.3 Dicofol (Kelthane)
5.3.4 Dieldrin
5.3.5 Endosulfan I, II
5.3.6 Endrin
5.3.7 Heptachlor Epoxide
5.3.8 Hexachlorobenzene
5.3.9 Lindane (y-hexachlorocyclohexane)
5.3.10 Mirex
5.3.11 Toxaphene
5.4 Organophosphate Pesticides
5.4.1 Chlorpyrifos
5.4.2 Diazinon
5.4.3 Disulfoton (Disyston)
5.4.4 Ethion
5.4.5 Terbufos
5.5 Chlorophenoxy Herbicides
5.5.1 Oxyfluorfen
5.6 Polycyclic Aromatic Hydrocarbons (PAHs)
5.6.1 Background
5.6.2 Pharmacokinetics
5.6.3 Acute Toxicity
5.6.4 Chronic Toxicity
5.6.5 Developmental Toxicity
5.6.6 Mutagenicity
5.6.7 Carcinogenicity
5.6.8 Special Susceptibilities
5.6.9 Interactive Effects
5.6.10 Critical Data Gaps
5.6.11 Summary of EPA Health Benchmarks
5.6.12 Major Sources
5.7 Polychlorinated Biphenyls (PCBs)
5.7.1 Background
5.7.2 Pharmacokinetics
5.7.3 Acute Toxicity
5.7.4 Chronic Toxicity
5.7.5 Developmental Toxicity
5.7.6 Mutagenicity
5.7.7 Carcinogenicity
5.7.8 Special Susceptibilities
5.7.9 Interactive Effects
Page
5-25
5-29
5-33
5-33
5-36
5-42
5-44
5-50
5-52
5-55
5-58
5-62
5-66
5-69
5-73
5-73
5-75
5-78
5-80
5-82
5-86
5-86
5-88
5-88
5-88
5-89
5-89
5-90
5-90
5-90
5-92
5-92
5-93
5-93
5-93
5-94
5-94
5-94
5-95
5-95
5-96
5-98
5-98
5-101
5-101
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TABLE OF CONTENTS
Section
Page
7
Appendix
A
B
C
D
E
F
5.7.10 Critical Data Gaps 5-101
5.7.11 Summary of EPA Health Benchmarks 5-101
5.7.12 Major Sources 5-101
5.8 Dioxins 5-102
5.8.1 Background 5-102
5.8.2 Pharmacokinetics 5-102
5.8.3 Acute Toxicity 5-103
5.8.4 Chronic Toxicity 5-103
5.8.5 Reproductive and Developmental Toxicity 5-104
5.8.6 Mutagenicity 5-104
5.8.7 Carcinogenicty 5-104
5.8.8 Special Susceptibilities 5-105
5.8.9 Interactive Effects 5-105
5.8.10 Critical Data Gaps 5-105
5.8.11 Summary of EPA Health Benchmarks 5-105
5.8.12 Major Sources 5-105
Mapping Tools for Risk Assessment and Risk Management 6-1
6.1 Overview of Population and Contaminant Mapping 6-1
6.2 Objectives of Mapping 6-1
6.3 Basic GIS Concepts for Population and
Contaminant Mapping 6-2
6.4 Internet Sources of Existing Data Files and GIS Coverages 6-5
6.5 Data Needed for Mapping 6-6
6.6 Mapping Programs 6-7
6.7 National Listing of Fish and Wildlife Advisories (NLFWA) Database . 6-8
Literature Cited 7-1
Reviewers of First Edition of Guidance Document A-1
Population Exposure Assessment—Consumption Patterns
and Surveys B-1
Dose Modification Due to Food Preparation and Cooking C-1
Guidance for Risk Characterization D-1
Additional Developmental toxicity Issues E-1
Summary of Limits of Detection for the Recommended
Target Analytes F-1
VI
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LIST OF FIGURES
LIST OF FIGURES
Number Page
1-1 Series Summary: Guidance for Assessing Chemical Contamination Data
for Use in Fish Advisories 1-11
2-1 Elements or Risk assessment and risk management 2-2
2-2 Schematic of exposure categories in upper half of a normal
population distribution 2-38
6-1 GIS Data Layers may use raster or vector
Representation techniques 6-3
6-2 Examples of GIS Displays from EPA's BASINInfo Maps-on-Demand
Facility 6-6
6-3 Map showing active fish and wildlife advisories for a state 6-7
VII
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LIST OF TABLES
LIST OF TABLES
Number Page
1-1 Target Analytes Recommended for Fish Sampling Programs 1-3
1-2 Comparison of FDA Action Levels and Tolerances with EPA Screening Values ... 1-6
1-3 Fish Consumption Rates for Various Fisher Populations 1-9
2-1 Uncertainty Factors and Modifying Factors for Estimating Exposure Limits for
Chronic Effects 2-17
2-2 Mean Body Weights of Children and Adults 2-29
2-3 Categories of Information Necessary for a Population Exposure Assessment . . . 2-34
2-4 Exposure Data Template 2-47
2-5 Risk Estimates 2-57
2-6 Risk Characterization 2-58
2-7 Risk Summaries for a Waterbody 2-59
2-8 Risk Summaries for a Geographic Area 2-61
3-1 Risk Values Used in Risk-Based Consumption Limit Tables 3-3
3-2 Input Parameters for Use in Risk Equations 3-6
3-3 Monthly Fish Consumption Limits for Carcinogenic and Noncarcinogenic Health
Endpoints—Chlordane 3-10
3-4 Monthly Consumption Limits for Carcinogenic and Noncarcinogenic Health
Endpoints- Chlordane 3-11
3-5 Average Body Weights and Associated Multipliers 3-13
4-1 Monthly Fish Consumption Limits for Carcinogenic and Noncarcinogenic
Health Endpoints - Arsenic 4-3
4-2 Monthly Fish Consumption Limits for Noncarcinogenic
Health Endpoint - Cadmium 4-4
4-3 Monthly Fish Consumption Limits for Noncarcinogenic
Health Endpoint - Methlymercury 4-5
4-4 Monthly Fish Consumption Limits for Noncarcinogenic
Health Endpoint - Selenium 4-6
4-5 Monthly Fish Consumption Limits for Noncarcinogenic
Health Endpoint - Tributyltin 4-7
4-6 Monthly Fish Consumption Limits for Carcinogenic and Noncarcinogenic
Health Endpoints - Chlordane 4-8
4-7 Monthly Fish Consumption Limits for Carcinogenic and Noncarcinogenic
Health Endpoints - DDT 4-9
4-8 Monthly Fish Consumption Limits for Noncarcinogenic
Health Endpoint - Dicofol 4-10
viii
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LIST OF TABLES
Number Page
4-9 Monthly Fish Consumption Limits for Carcinogenic and Noncarcinogenic
Health Endpoints -Dieldrin 4-11
4-10 Monthly Fish Consumption Limits for Noncarcinogenic
Health Endpoint - Endosulfan 4-12
4-11 Monthly Fish Consumption Limits for Noncarcinogenic
Health Endpoint - Endrin 4-13
4-12 Monthly Fish Consumption Limits for Carcinogenic and Noncarcinogenic
Health Endpoints - Heptachlor Epoxide 4-14
4-13 Monthly Fish Consumption Limits for Carcinogenic and Noncarcinogenic
Health Endpoints - Hexachlorobenzene 4-15
4-14 Monthly Fish Consumption Limits for Carcinogenic and Noncarcinogenic
Health Endpoints - Lindane 4-16
4-15 Monthly Fish Consumption Limits for Carcinogenic and Noncarcinogenic
Health Endpoints - Mirex 4-17
4-16 Monthly Fish Consumption Limits for Carcinogenic and Noncarcinogenic
Health Endpoints - Toxaphene 4-18
4-17 Monthly Fish Consumption Limits for Noncarcinogenic
Health Endpoint - Chlorpyrifos 4-19
4-18 Monthly Fish Consumption Limits for Noncarcinogenic
Health Endpoint - Diazinon 4-20
4-19 Monthly Fish Consumption Limits for Noncarcinogenic
Health Endpoint - Disulfoton 4-21
4-20 Monthly Fish Consumption Limits for Noncarcinogenic
Health Endpoint - Ethion 4-22
4-21 Monthly Fish Consumption Limits for Noncarcinogenic
Health Endpoint - Terbufos 4-23
4-22 Monthly Fish Consumption Limits for Noncarcinogenic
Health Endpoint - Oxyfluorfen 4-24
4-23 Monthly Fish Consumption Limits for Carcinogenic
Health Endpoint - PAHs 4-25
4-24 Monthly Fish Consumption Limits for Carcinogenic and Noncarcinogenic
Health Endpoints - PCBs 4-26
4-25 Monthly Fish Consumption Limits for Carcinogenic
Health Endpoint - Dioxins/Furans 4-27
5-1 Health and Toxicological Data Reviewed for Target Analytes 5-2
5-2 Toxicity Equivalent Factors for Various PAHs 5-91
5-3 Relative Potency Estimates for Various PAHs 5-92
5-4 Reported Concentrations (ppm) of Dioxin-Like Congeners in Commercial
Aroclor Mixtures 5-100
5-5 PCB and Dioxin Concentrations (ppb) in Channel Catfish 5-100
6-1 Comparison of Raster- Versus Vector-Based GIS Programs 6-4
IX
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LIST OF ACRONYMS
LIST OF ACRONYMS
ACTH adrenocortical trophic hormone
ARL acceptable lifetime risk level
ATSDR Agency for Toxic Substances and Disease Registry
BCF bioconcentration factor
BW body weight
CAG Carcinogenic Assessment Group
CCRIS chemical carcinogenesis Research Information System
CDDs chlorodibenzo-p-dioxins
CDF chlorodibenzofurans
CERCLA Comprehensive Environmental Response, Compensation,
and Liability Act
CERCLIS CERCLA List of Sites
CMS central nervous system
COC chain-of-custody
CR consumption rate
CSF cancer slope factor
ODD p,p1-dichlorodiphenyldichloroethane
DDE p,p1-dichlorodiphenyldichloroethylene
DDT p,p1-dichlorodiphenyltrichloroethane
EPA U.S. Environmental Protection Agency
FDA U.S. Food and Drug Administration
FGDC Federal Geographic Data Committee
FIFRA Federal Insecticide, Fungicide, and Rodenticide Act
y-BHC benzene hexachloride
y-HCH hexachlorocyclohexane
GC/ECD gaschromatography/electron capture detection
GC/MS gaschromatography/mass spectrometry
Gl gastrointestinal
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LIST OF ACRONYMS
GIS
GPS
HEAST
HRGC/HRMS
HSDB
IRIS
LD50
LEL
Lit
LMS
LOAEL
LOD
MF
MFC
MOE
MS
MAS
NFTDR
NGO
NHANESII
NIOSH
NLFWA
NOAA
NOAEL
NSCRF
NSDI
NTP
OAPCA
OPP
PAHs
PCBs
PCDDs
geographic information system
Global Positioning System
Health Effects Assessment Summary Tables
high-resolution gas chromatography/high-resolution mass
spectrometry
Hazardous Substances Data Bank
Integrated Risk Information System
lethal dose, 50% kill
lowest exposure limit
luteinizing hormone
linearized multistage (model)
lowest observed adverse effects level
limit of detection
modifying factor
mixed function oxidase
margin of exposure
meal size
National Academy of Sciences
National Fish Tissue Data Repository
nongovernmental organization
National Health and Nutrition Examination Survey
National Institute of Occupational Safety and Health
National Listing of Fish and Wildlife Advisories
National Oceanic and Atmospheric Administration
no observable adverse effect level
National Study of Chemical Residues in Fish
National Spatial Data Infrastructure
National Toxicology Program
Organotin Antifouling Paint Control Act
Office of Pesticide Programs
polycyclic aromatic hydrocarbons
polychlorinated biphenyls
polychlorinated dibenzo-p-dioxins
XI
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LIST OF ACRONYMS
PCDFs
PCS
PEC
PNAs
POTW
QA
QC
RAC
RBC
RCS
RDA
RfD
RTECs
SAB
SCE
SVs
2,4,5-T
2,3,7,8-TCDD
2,3,7,8-TCDF
2,4,5-TCP
TEC
TRI
UF
USDA
USFWS
USGS
WHO
WOE
polychlorinated dibenzofurans
Permit Compliance System
potency equivalency concentration
polynuclear aromatic hydrocarbons
publically owned treatment works
quality assurance
quality control
reference ambient concentrations
red blood cell
Relative Source Contribution
recommended dietary allowance
reference dose
Registry of Toxic Effects of Chemical Substances
Science Advisory Board
sister chromatid exchange
screening values
2,4,5-trichlorophenoxyaceticacid
2,3,7,8-tetrachlorodibenzo-p-dioxin
2,3,7,8-tetrachlorodibenzofuran
2,4,5-trichlorophenol
toxicity equivalent concentrations
Toxic Release Inventory
uncertainty factor
U.S. Department of Agriculture
U.S. Fish and Wildlife Service
U.S. Geological Survey
World Health Organization
weight of evidence
XII
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ACKNOWLEDGMENTS
ACKNOWLEDGMENTS
This report was prepared by the U.S. Environmental Protection Agency, Office of
Water, Fish Contaminant Section. The EPA Work Assignment Manager for this
document was Jeffrey Bigler who provided overall project coordination as well as
technical direction. EPA was supported in the development of the third edition of
this document by the Research Triangle Institute (RTI), Inc. (EPA Contract No.
68-C7-0056). Patricia Cunningham and Susan Goldhaber of RTI prepared this
third edition. EPA was supported in the development of the original document by
Abt Associates and Tetra Tech, Inc. Kathleen Cunningham of Abt Associates was
the contractor's Project Manager. Preparation of the first edition of this guidance
document was facilitated by the substantial efforts of the numerous Workgroup
members and reviewers (see Appendix A).
XIII
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EXECUTIVE SUMMARY
EXECUTIVE SUMMARY
State, local, tribal, and federal agencies currently use various methods to estimate
risks to human health from the consumption of chemically contaminated,
noncommercially caught fish and shellfish. A 1988 survey, funded by the U.S.
Environmental Protection Agency (EPA) and conducted by the American Fisheries
Society, identified the need for standardizing the approaches to evaluating risks
and developing fish consumption advisories that are comparable across different
jurisdictions. Four key components were identified as critical to the development
of a consistent risk-based approach: standardized practices for sampling and
analyzing fish, standardized risk assessment methods, standardized procedures
for making risk management decisions, and standardized approaches to risk
communication.
To address concerns raised by the survey respondents, EPA has developed a
series of four documents designed to provide guidance to state, local, tribal, and
regional environmental health officials responsible for issuing fish consumption
advisories. The documents are meant to provide guidance only and do not
constitute a regulatory requirement. The documents are:
Guidance for Assessing Chemical Contaminant Data for Use in Fish Advisories
Volume 1: Fish Sampling and Analysis
Volume 2: Risk Assessment and Fish Consumption Limits
Volume 3: Overview of Risk Management
Volume 4: Risk Communication.
Volume 1 was first released in September 1993, and a second edition followed in
September 1995. Volume 2 was first released in June 1994 and was followed by
a second edition in July 1997. Volume 3 was released in June 1996, and Volume
4 was released in March 1995. It is essential that all four documents be used
together, since no single volume addresses all of the topics involved in the
development of risk-based fish consumption advisories.
The objective of Volume 2: Risk Assessment and Fish Consumption Limits is to
provide guidance on the development of risk-based meal consumption limits for
25 high-priority chemical contaminants (target analytes). The target analytes
addressed in this guidance series were selected by EPA's Office of Water as
particularly significant contaminants, based on their documented occurrence in
fish and shellfish, their persistence in the environment, their potential for
bioaccumulation, and their oral toxicity to humans. The criteria for their selection
are discussed in Section 4 of Volume 1 of this series.
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EXECUTIVE SUMMARY
In addition to presenting monthly consumption limit tables, Volume 2 discusses
risk assessment methods used to derive the limits and discusses procedures
used to modify these limits to reflect local conditions. A toxicological profile
summary for each of the target analytes presenting current toxicity data is also
provided. Additional sources of information are listed for those seeking a more in-
depth discussion of risk assessment methods.
The first edition of Volume 2 was reviewed by experts at the federal, state, tribal,
and local levels who were members of the Fish Contaminant Workgroup. These
individuals contributed significant technical information and guidance during the
development of this document. Their input was used to revise the document to
make it more useful and informative to public health professionals. The workgroup
was not involved in reviewing this third edition because the basic risk assessment
procedures had already been approved. This third edition was issued to update
toxicological information for several of the target analytes; to incorporate the
Agency's new health risk information, daily consumption rates, and body weight
assumptions into the body of the document; and to reformat the monthly
consumption limit tables.
This third edition provides risk assessors and managers with the most current
toxicological information for each of the 25 target analytes and provides users with
• Detailed information on risk assessment methods, including information on
population exposure, fish consumption patterns, consumption surveys, risk
reduction through the use of various preparation and cooking procedures, and
risk characterization (Section 2)
• Reformatted monthly consumption limits tables and instructions on how these
tables can be modified to reflect local site-specific conditions for specific
populations of concern (Section 3, Section 4)
• A toxicological profile summarizing current toxicity data for each target analyte
(Section 5)
• A brief explanation of geographic information system (GIS) mapping tools for
use in risk assessment and risk management (Section 6).
The information in this document may be used in conjunction with contaminant
data from local fish and shellfish sampling programs and fish consumption
surveys (or from fish consumption data provided in Appendix D), to select or
calculate risk-based consumption limits for contaminated noncommercially caught
fish and shellfish. The consumption limits may be used with other types of
information (e.g., cultural and dietary characteristics of the populations of concern,
social and economic impacts, and health issues, including benefits of fish
consumption and accessibility of other food sources) to establish health
advisories. The decision-making process for the development of fish advisories
is discussed in the risk management document in this series (Volume 3).
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EXECUTIVE SUMMARY
EPA welcomes your suggestions and comments. A major goal of this guidance
document series is to provide a clear and usable summary of critical information
necessary to make informed decisions concerning fish consumption advisories.
We encourage comments and hope this document will be a useful adjunct to the
resources used by states, local governments, and tribal organizations in making
decisions concerning fish advisories.
XVI
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1. INTRODUCTION
SECTION 1
INTRODUCTION
1.1 OVERVIEW
Toxic chemicals released to the environment from point sources such as industrial
and municipal discharges and from nonpoint sources such as agricultural runoff
and atmospheric deposition have contaminated surface waters and their
sediments across the United States. In some areas, contamination arises from
one or more related chemicals. For example, in the Hudson River in New York,
attention has focused on high concentrations of a group of related chemicals
called polychlorinated biphenyls, or PCBs. In other areas, a complex mixture of
chemicals is present. For example, over 900 different synthetic organic
compounds have been found in Puget Sound in Washington State, while nearly
1,000 chemical contaminants have reportedly been found in the Great Lakes.
Many chemical pollutants concentrate in fish and shellfish by accumulating in fatty
tissues or selectively binding to fish muscle tissue (the fillet). Even extremely low
concentrations of bioaccumulative pollutants detected in water or bottom
sediments may result in fish or shellfish tissue concentrations high enough to pose
health risks to fish consumers. Lipophilic contaminants, particularly certain
organochlorine compounds, tend to accumulate in the fatty tissues of fish.
Consequently, fish species with a higher fat content, such as carp, bluefish, some
species of salmon, and catfish, may pose greater risks from some contaminants
than leaner fish such as bass, sunfish, and yellow perch. Although exposure to
some contaminants may be reduced by removing the fat, skin, and viscera before
the fish is eaten, other contaminants, such as methylmercury, accumulate in the
muscle tissue of the fillet and therefore cannot be removed by trimming. In
addition, some fish are consumed whole or are used whole in the preparation of
fish stock for soups and other foods. Under these conditions, the entire body
burden of bioaccumulative contaminants contained in the fish would be ingested
by the consumer (U.S. EPA, 1991b).
Results of a 1989 survey of methods to estimate risks to human health from
consumption of chemically contaminated fish (Cunningham et al.,1990), funded
by the U.S. Environmental Protection Agency (EPA) and conducted by the
American Fisheries Society, identified the need for standardizing the approaches
to assessing risks and for developing advisories for contaminated fish and
shellfish. Four key components were identified as critical to the development of
a consistent risk-based approach to developing consumption advisories: standard
1-1
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1. INTRODUCTION
practices for sampling and analyzing fish and shellfish, standardized risk
assessment methods, standardized procedures for making risk management
decisions, and standardized approaches to risk communication.
Note: Throughout this document series, the term "fish" refers to sport-
and subsistence-caught freshwater, estuarine, and marine fish and
shellfish, unless otherwise noted.
To address concerns raised by the survey, EPA developed a series of four
documents designed to provide guidance to state, local, regional, and tribal
environmental health officials who are responsible for issuing fish consumption
advisories for noncommercially caught fish. The documents are meant to provide
guidance only and do not constitute a regulatory requirement. The documents are:
Guidance for Assessing Chemical Contamination Data for Use in Fish Advisories,
Volume 1: Fish Sampling and Analysis (released 1993, revised in 1995 and 2000),
Volume 2: Risk Assessment and Fish Consumption Limits (released in 1994 and
revised in 1997 and 2000), Volume 3: Risk Management (released in 1996), and
Volume 4: Risk Communication (released in 1995). EPA recommends that the
four volumes of this guidance series be used together, since no one volume
provides all the necessary information to make decisions regarding the issuance
of fish consumption advisories.
This volume (Volume 2) provides guidance on risk assessment procedures to use
in the development of risk-based consumption limits for the 25 high-priority chem-
ical contaminants identified in Volume 1 (see Table 1-1).
The target analytes listed in Table 1-1 were selected by EPA's Office of Water as
particularly significant fish contaminants, based on their occurrence in fish and
shellfish (as evidenced by their detection in regional or national fish monitoring
programs or by state issuance of a fish advisory), their persistence in the
environment (half-life >30 days), their potential for bioaccumulation (BCF values
>300), and their oral toxicity to humans.
1.2 OBJECTIVES
It should be noted that the EPA methodology described in both Volumes 1 and 2
of this guidance series offers great flexibility to the state users. These documents
are designed to meet the objectives of state monitoring and risk assessment
programs by providing options to meet specific state or study needs within state
budgetary constraints. The users of this fish advisory guidance document should
recognize that it is the consistent application of the EPA methodology and
processes rather than individual elements of the program sampling design that
are of major importance in improving consistency among state fish advisory
1-2
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1. INTRODUCTION
Table 1-1. Target Analytes Recommended for Fish Sampling Programs
Metals
Arsenic (inorganic)
Cadmium
Mercury (methylmercury)
Selenium
Tributyltin
Organochlorine Pesticides
Chlordane, total (c/s- and frans-chlordane,
c/s- and frans-nonachlor, oxychlordane)
DDT, total (2,4-DDD, 4,4'-DDD, 2,4'-DDE,
4,4-DDE, 2,4-DDT, 4,4'-DDT)
Dicofol
Dieldrin
Endosulfan (I and II)
Endrin
Heptachlor epoxide"
Hexachlorobenzene
Lindane (y-hexachlorocyclohexane; y-HCH)0
Mirexd
Toxaphene
Organophosphate Pesticides
Chlorpyrifos
Diazinon
Disulfoton
Ethion
Terbufos
Chlorophenoxy Herbicides
Oxyfluorfen
PAHse
PCBs
Total PCBs (sum of PCB congeners or
Aroclors)'
Dioxins/furans9
ODD = p,p' - dichlorodiphenyldichloroethane.
DDE = p,p' - dichlorodiphenyldichloroethylene.
DDT = p,p' - dichlorodiphenyltrichloroethane.
PAHs = Polycyclic aromatic hydrocarbons.
PCBs = Polychlorinated biphenyls.
a The reader should note that carbophenothion was included on the original list of target analytes. Because the
registrant did not support reregistration of this chemical, all registered uses were canceled after December
1989. For this reason and because of its use profile, carbophenothion was removed from the recommended list
of target analytes.
b Heptachlor epoxide is not a pesticide but is a metabolite of the pesticide heptachlor.
c Also known as y-benzene hexachloride (y-BHC).
d Mirex should be regarded primarily as a regional target analyte in the southeast and Great Lakes states, unless
historic tissue, sediment, or discharge data indicate the likelihood of its presence in other areas.
e It is recommended that tissue samples be analyzed for benzo[a]pyrene and 14 other PAHs and that the order-
of-magnitude relative potencies given for these PAHs be used to calculate a potency equivalency concentration
(PEC) for each sample (see Section 5 of Volume 1).
f Analysis of total PCBs (as the sum of Aroclors or PCB congeners) is recommended for conducting human
health risk assessments for total PCBs (see Sections 4.3.6 and 5.3.2.6 of Volume 1). A standard method for
Aroclor analysis is available (EPA Method 608). A standard method for congener analyses is under
development by EPA; however, it has not been finalized. States that currently do congener-specific PCB
analyses should continue to do so and other states are encouraged to develop the capability to conduct PCB
congener analyses. When standard methods for congener analysis have been verified and peer-reviewed, the
Office of Water will evaluate the use of these methods.
9 It is recommended that the 17 2,3,7,8-substituted tetra-through octa-chlorinated dibenzo-p-dioxins (PCDDs)
and dibenzofurans (PCDFs) and the 12 dioxin-like PCBs be determined and a toxicity-weighted total
concentration calculated for each sample (Van den Berg et al., 1998) (see Sections 4.3.7 and 5.3.2.6 of
Volume 1).
1-3
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1. INTRODUCTION
programs. For example, this document presents consumption limits that were
calculated using a risk level of 1 in 100,000 (10~5); however, states may choose
to calculate consumption limits based on other risk levels.
One major factor currently affecting the comparability of fish advisory information
nationwide is the fact that the states employ different methodologies to determine
the necessity for issuing an advisory. For example, some states currently do not
use the EPA methodology at all or use it only in their assessment of health risks
for certain chemical contaminants. Often these states rely instead on
exceedances of U.S. Food and Drug Administration (FDA) action levels or
tolerances to determine the need to issue an advisory. FDA's mission is to
protect the public health with respect to levels of chemical contaminants in all
foods, including fish and shellfish.
FDA has developed both action levels and tolerances to address levels of
contamination in foods. FDA may establish an action level when food contains a
chemical from sources of contamination that cannot be avoided even by
adherence to good agricultural or manufacturing practices, such as contamination
by a pesticide that persists in the environment. An action level is an administrative
guideline or instruction to the agency field unit that defines the extent of
contamination at which FDA may regard food as adulterated. An action level
represents the limit at or above which FDA may take legal action to remove
products from the marketplace. Under the Food, Drug, and Cosmetic Act, FDA
also may set tolerances for unavoidably added poisonous or deleterious
substances, that is, substances that are either required in the production of food
or are otherwise unavoidable by good manufacturing practices. A tolerance is a
regulation that is established following formal, rulemaking procedures; an action
level is a guideline or "instruction" and is not a formal regulation (Boyer et al.,
1991).
FDA's jurisdiction in setting action levels or tolerances is limited to contaminants
in food shipped and marketed in interstate commerce. Thus, the methodology
used by FDA in establishing action levels or tolerances is to determine the health
risks of chemical contaminants in fish and shellfish that are bought and sold in
interstate commerce rather than in locally harvested fish and shellfish (Bolger et
al., 1990). FDA action levels and tolerances are indicators of chemical residue
levels in fish and shellfish that should not be exceeded for the general population
who consume fish and shellfish typically purchased in supermarkets or fish
markets that sell products that are harvested from a wide geographic area,
including imported fish and shellfish products. However, the underlying
assumptions used in the FDA methodology were never intended to be protective
of recreational, tribal, ethnic, and subsistence fishers who typically consume larger
quantities of fish than the general population and often harvest the fish and
shellfish they consume from the same local waterbodies repeatedly over many
years. If these local fishing and harvesting areas contain fish and shellfish with
elevated tissue levels of chemical contaminants, these individuals potentially
1-4
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1. INTRODUCTION
could have increased health risks associated with their consumption of fish and
shellfish.
The following chemical contaminants discussed in this volume have FDA action
levels for their concentration in the edible portion offish and shellfish: chlordane,
DDT, DDE, ODD, dieldrin, heptachlor epoxide, mercury, and mirex. FDA has not
set an action level for PCBs in fish, but has established a tolerance in fish for this
chemical. FDA also has set action levels in fish for two chemical contaminants
that are not discussed in this volume: chlordecone (Kepone) and ethylene
dibromide. FDA had set an action level for toxaphene; however, this level was
revoked in 1993 because FDA determined that toxaphene residues were no
longer occurring as unavoidable contaminants in food (57 FR 60859). In addition,
in 1981, FDA set an advisory level for dioxin in fish, in response to requests from
the governors of the Great Lake states. This advisory level was nonenforceable
federal advice and was provided with the intention that state and local authorities
use it to develop their own control policies (Boyer et al., 1990).
Table 1-2 compares the FDA action levels and tolerances for these seven
chemical contaminants with EPA's recommended screening values (SVs) for
recreational and subsistence fishers calculated for these target analytes using the
EPA methodology.
The EPA SV for each chemical contaminant is defined as the concentration of the
chemical in fish tissue that is of potential public health concern and that is used
as a threshold value against which tissue residue levels of the contaminant in fish
and shellfish can be compared. The SV is calculated based on both the
noncarcinogenic and carcinogenic effects of the chemical contaminant, which are
discussed in detail in Volume 1 of this series (EPA, 2000a). EPA recommends
that the more conservative of the calculated values derived from the
noncarcinogenic rather than the carcinogenic effects be used because it is more
protective of the consumer population (either recreational or subsistence fishers).
As can be seen in Table 1-2, for the recreational fisher, the EPA-recommended
values typically range from 2 to 120 times lower and thus are more protective than
the corresponding FDA action or tolerance level. This difference is even more
striking for subsistence fishers for whom the SVs are 20 to 977 times lower than
the FDA values.
EPA and FDA have agreed that the use of FDA action levels for the purposes of
making local advisory determinations is inappropriate. In letters to all states,
guidance documents, and annual conferences, this practice has been
discouraged by EPA and FDA in favor of EPA's risk-based approach to derive
local fish consumption advisories.
1-5
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1. INTRODUCTION
Table 1-2. Comparison of FDA Action Levels and Tolerances with EPA
Screening Values
Chemical
Contaminant
Chlordane
Total DDT
Dieldrin
Heptachlor epoxide
Mercury
Mi rex
PCBs
FDA Action
Level (ppm)
0.3
5.0
0.3
0.3
1.0
0.1
FDA Tolerance
Level (ppm)
2.0
EPA SV for
Recreational
Fishers (ppm)
0.114
0.117
2.5 x10'3
4.39 x10'3
0.40
0.80
0.02
EPA SV for
Subsistence
Fishers (ppm)
0.014
0.014
3.07x10'4
5.40 x10'4
0.049
0.098
2.45 x10'3
Source: U.S. FDA, 1998.
1.3 SENSITIVE SUBPOPULATIONS
In addition to the risks borne by the general population as a result of consuming
contaminated fish, various populations eating higher-than-average quantities of
fish are at greater risk of having higher body burdens of bioaccumulative
contaminants. Those at greatest risk include sport and subsistence fishers. In this
document, subsistence fishers are defined as fishers who rely on noncommer-
cially caught fish and shellfish as a major source of protein in their diets. In
addition to these populations, pregnant women and children may be at greater
risk of incurring adverse effects than other members of the populations because
of their proportionally higher consumption rates and/or increased susceptibility to
adverse toxicological effects.
EPA has provided this guidance to be especially protective of recreational fishers
and subsistence fishers within the general U.S. population. EPA recognizes,
however, that Native American subsistence fishers are a unique subsistence
fisher population that needs to be considered separately. For Native American
subsistence fishers, eating fish is not simply a dietary choice that can be
completely eliminated if chemical contamination reaches unacceptable levels;
rather eating fish is an integral part of their lifestyle and culture. This traditional
lifestyle is a living religion that includes values about environmental responsibility
and community health as taught by elders and tribal religious leaders (Harris and
Harper, 1997). Therefore, methods for balancing benefits and risks from eating
1-6
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1. INTRODUCTION
contaminated fish must be evaluated differently than for the general fisher
population.
For any given population, there can be a sensitive subpopulation comprising
individuals who may be at higher than average risk due to their increased
exposure or their increased sensitivity to a contaminant or both. For Native
American subsistence fishers, exposure issues of concern that should be
addressed as part of a comprehensive exposure assessment include the
following:
• Consumption rates and dietary preferences. Harris and Harper (1997)
surveyed traditional tribal members in Oregon with a subsistence lifestyle and
determined a consumption rate of 540 g/d that included fresh, dried, and
smoked fish. They also confirmed that the parts of the fish (heads, fins,
skeleton, and eggs) that were eaten by this group were not typically eaten by
other groups. Another study conducted of four tribes in the Northwest that
also surveyed tribal members in Oregon, but did not target subsistence
fishers, reported a 99th percentile ingestion rate of 390 g/d for tribal members
(CRITFC, 1994). These consumption rates are much higher than the default
consumption rates provided in this document for subsistence fishers, which
emphasizes the need to identify the consumption rate of the Native American
subsistence population of concern.
• Community characteristics. It is important to consider family-specific
fishing patterns in any exposure scenario, and attention should be paid to the
role of the fishing family with respect to the tribal distribution of fish, the
sharing ethic, and providing fish for ceremonial/religious events. Entire
communities are exposed if fish are contaminated, and the community
contaminant burden as a whole must be considered, not just the maximally
exposed individual.
• Multiple contaminant exposures. Multiple contaminant exposure is
significant for Native American subsistence fishers. A large number of
contaminants are often detected in fish tissues and their combined risk
associated with the higher consumption rates and dietary preferences for
certain fish parts could be very high even if individual contaminants do not
exceed the EPA reference dose (Harper and Harris, 1999).
• Other exposure pathways. For Native American subsistence fishers, overall
exposure to a contaminant may be underestimated if it fails to take into
account nonfood uses of fish and other animal parts that may contribute to
overall exposure, such as using teeth and bones for decorations and whistles,
animal skins for clothing, and rendered fish belly fat for body paint (Harper
and Harris, 1999). If other wildlife species (e.g., feral mammals, turtles,
waterfowl) that also live in or drink from the contaminated waterbody are
eaten, or if the contaminated water is used for irrigation of crops or for
livestock watering or human drinking water, the relative source contribution
1-7
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1. INTRODUCTION
of these other pathways of exposure also must be considered. As with fish
and wild game, plants are used by Native Americans for more than just
nutrition. Daily cleaning, preparation, and consumption of plants and crafting
of plant materials into household goods occurs throughout the year (Harris
and Harper, 1997).
As in the general population, increased sensitivity to a chemical contaminant for
Native Americans can result from factors such as an individual's underlying health
status and medications, baseline dietary composition and quality, genetics,
socioeconomic status, access to health care, quality of replacement protein, age,
gender, pregnancy, and lactation. These factors are only partially considered in
the uncertainty factor(s) used to develop an RfD (Harper and Harris, 1999).
Other important issues that need to be considered concern risk characterization
and risk management. For Native American subsistence fishers, the use of an
acceptable risk level of 1 in 100,000 (10~5) may not be acceptable to all tribes.
Each tribe has the right to decide for themselves what an acceptable level of risk
is, and, in some cases, it may be zero risk to protect cultural resources.
Ecological well-being or health is another key issue. Human health and ecological
health are connected in many ways, and the ripple effects are often not
recognized. For example, human health may be affected by injury to the
environment, which affects the economy and the culture (Harper and Harris,
1999).
Native American subsistence fishers should be treated as a special high-risk
group of fish consumers distinct from fishers in the general population and distinct
even from other Native American fish consumers living in more suburbanized
communities. Table 1-3 compares fish consumption rates for various fisher
populations within the general population and specific Native American tribal
populations. EPA currently recommends default fish consumption rates of 17.5
g/d for recreational fishers and 142.4 g/d for subsistence fishers. However, the
tribal population fish consumption studies show that some Native American tribal
members living in river-based communities (CRITFC, 1994) eat from 3 to 22 times
more fish (from 59 g/d up to 390 g/d) than recreational fishers, and that traditional
Native American subsistence fishing families may eat up to 30 times more fish,
almost 1.2 1b/d (540 g/d) (Harris and Harper 1997). The fish consumption rate
from Harris and Harper (1997) for Native American subsistence fishers (540 g/d)
is also 3.8 times higher than the EPA default consumption rate for subsistence
fishers (142.4 g/d) in the general population. The difference in fish consumption
is due to the fact that the Native American subsistence fisher's lifestyle is not the
same as a recreational fisher's lifestyle with additional fish consumption added,
nor is it the same as the "average" Native American tribal member living in a fairly
suburbanized tribal community. In addition to exposures from direct consumption
of contaminated fish, Native American subsistence fishers also receive more
exposure to the water and sediments associated with catching and preparing fish,
1-8
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Table 1-3. Fish Consumption Rates for Various Fisher Populations
Source
U.S. EPA
Harris and
Harper (1997)
CRITFC
(1994)
Toy et al.
(1996)
Recreational
fishers (g/d)
17. 5a
NA
NA
NA
Subsistence
Fishers (g/d)
142.4 a
NA
NA
NA
Native American
Subsistence
fishers (g/d)
70 (mean) b
170 (95th
percentile)b
540 (fresh, smoked,
and dried)
NA
NA
Native Americans (g/d)
NA
NA
59 (mean)
170 (95th percentile)
390 (99th percentile)
53 (median, males)
34 (median, females)
66 (median, males)
25 (median, females)
Basis for Consumption Rate
Fish consumption rate from 1994
and 1996 Continuing Survey of
Food Intake by Individuals
(CSFII) (USDA/ARS, 1998)
Surveyed members of the
Confederated Tribes of the
Umatilla Indian Reservation
Surveyed members of the
Umatilla, Nez Perce, Yakama,
and Warm Springs Tribes
Surveyed members of the
Tulalip Tribe
Surveyed members of the
Squaxin Island Tribe
NA = Not available.
a These values were revised in the 3rd edition of Volume 1 of this series (U.S. EPA, 2000a)
b These values are from EPA's Exposure Factors Handbook (U.S. EPA, 1997f)
O
O
c
O
CD
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1. INTRODUCTION
and possibly from drinking more unfiltered river water than more suburbanized
tribal community members as well. The Native American subsistence fishing
population should be treated as a separate group with a very unique lifestyle,
distinct from recreational and subsistence fishers in the general U.S. population
and even distinct from other Native American fisher populations.
1.4 CONTENTS OF VOLUME 2
Figure 1-1 shows how Volume 2 fits into the overall guidance series and lists the
major categories of information provided. This volume covers topics necessary for
conducting risk assessments related to consumption of chemically contaminated
fish. The first four sections follow the anticipated sequence of activities to conduct
a risk assessment, develop risk-based consumption limits, and prepare consump-
tion limit tables for a range of fish contaminant levels, meal sizes, and consumer
groups. The last two sections provide summary information on the toxicological
properties of the 25 target analytes and geographic information system (GIS)
mapping tools for risk assessment and risk management.
Section 1 of this document reviews the development of this guidance document
series, lists the 25 target analytes of concern with respect to chemical contamina-
tion offish and shellfish, summarizes additions and revisions to this third edition,
and references information used in the development of this document.
Section 2 introduces the EPA four-step risk assessment process: hazard identifi-
cation, dose-response assessment, exposure assessment, and risk characteriza-
tion. Details on each of these steps are provided, along with a discussion of the
major uncertainties and assumptions.
Section 3 of this document presents the information needed to calculate or modify
the consumption limit tables provided for the 25 target analytes in Section 4. The
reader is guided through calculations of risk-based consumption limits for
carcinogenic and noncarcinogenic effects using the appropriate cancer slope
factor (CSF) and reference dose (RfD). The reader is shown how selection of
various input parameters such as the maximum acceptable risk level, consumer
body weight, meal size, and time-averaging period influence fish consumption
limits for single species diets. In addition, information is provided on methods for
calculating consumption limits for single-species diets with multiple contaminants
and multiple-species diets contaminated with a single or multiple contaminants.
The monthly consumption limits for each of the 25 target analytes are provided
in Section 4.
Section 5 presents a toxicological profile summary for each of the 25 target
analytes. Each profile summary contains a discussion of the pharmacokinetics,
acute toxicity, chronic toxicity, reproductive and developmental toxicity,
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1. INTRODUCTION
Volume 1: Fish
Sampling and Analysis
Volume 2: Risk
Assessment and Fish
Consumption Limits
Volume 3: Overview
of Risk Management
Volume 4: Risk
Communication
1. Introduction
2. Risk Assessment
Methods
3. Development and Use of
Risk-based Consumption
Limits
4. Risk-based Consumption
Limit Tables
5. Toxicological Profile
Summaries for Target
Analytes
6. Mapping Tools for Risk
Assessment and Risk
Management
7. Literature Cited
Figure 1-1. Series Summary: Guidancefor Assessing Chemical
Contaminant Data for Use in Fish Advisories.
1-11
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1. INTRODUCTION
mutagenicity, carcinogenicity, populations with special susceptibilities, interactive
effects of the target analytes with other chemical contaminants, and critical data
gaps with respect to toxicity. The most current EPA risk values (CSFs and RfDs)
from sources such as EPA's Integrated Risk Information System (IRIS) and the
Office of Pesticide Programs are provided, with a discussion of supporting dose-
response data.
Section 6 has been added to provide readers with an overview of GIS mapping
tools for use in risk assessment and risk management. Mapping can be used to
display information germane to all aspects offish advisory programs. Maps may
focus on fish contaminant levels, waterbodies where fish advisories are in effect,
sport and subsistence fishing locations, or consumption levels of target popula-
tions of fishers. The reader is provided with instructions to access EPA websites
on the Internet to obtain additional GIS datasets and coverages.
In keeping with current EPA recommendations, discussions of uncertainty and
assumptions are included in each section of the document. Although information
was sought from a variety of sources to provide the best available data
concerning the development of fish consumption advisories, limited data exist for
some critical parameters (e.g., toxicological properties of certain chemicals and
susceptibilities of specific populations such as the elderly, children, and pregnant
or nursing women). Although substantial toxicological information is available for
all target analytes discussed in this document, readers are cautioned to always
consider the methods and values presented in the context of the uncertainty
inherent in the application of science to policies for safeguarding the general
public from environmental hazards.
The focus of this document is primarily on the risk due to consumption of non-
commercially caught fish and shellfish from freshwater, estuarine, and marine
waters. This document provides guidance on the evaluation of the overall risk
associated with multimedia exposure to chemical contaminants found in fish (e.g.,
exposure resulting from other food sources, consumer products, air, water, and
soil). EPA recommends that a comprehensive risk assessment be considered for
all confirmed fish contaminants, including an evaluation of all significant exposure
pathways (e.g., inhalation, dermal, and oral exposures).
Risk assessment and risk management of chemically contaminated fish are
complex processes because of the many considerations involved in setting fish
consumption advisories, including both the health risks and benefits of fish
consumption, the roles of state and federal agencies, and the potential impact of
advisories on economic and societal factors. These topics are discussed in
Volume 3 of this guidance series (Overview of Risk Management). The final
volume in the series deals with how risk managers can best communicate the
health risks and benefits of fish consumption to the general public as well as
recreational and subsistence fishers. These topics are detailed in Volume 4 (Risk
Communication).
1-12
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1. INTRODUCTION
1.5 CHANGES TO VOLUME 2
The following changes were made to this edition:
Section 1:
• Included discussion of Native American subsistence fishers.
• Included new information on the development of FDA action levels and
tolerances and provided rationale as to why states should adopt the EPA risk-
based approach.
Section 2:
• Revised table on uncertainty factors to be consistent with new information.
• Revised developmental toxicity section: removed repetitive material and put
detailed information from this section in Appendix E.
• Included information from recent EPA guidelines for the health risk
assessment of chemical mixtures (1999).
Section 3:
• Revised consumption limit tables in Section 4 to be calculated as fish meals
per month, at various fish tissue concentrations, for noncancer and cancer
health endpoints.
• Assumed an acceptable risk of 1 in 100,000 in meal consumption limits; the
second edition used an acceptable risk of 1 in 10,000, 1 in 100,000, and 1 in
a million.
• Updated risk values used in consumption limit tables based on IRIS (1999)
and new information from EPA's Office of Pesticide Programs.
• Assumed an 8-oz (0.227-kg) meal size for calculation consumption limits; the
second edition assumed four meal sizes of 4, 8, 12, and 16 oz.
• Recommended a default value for meal size of shellfish.
• Assumed a monthly time-averaging period; the second edition assumed
biweekly, 10-day, weekly, and monthly time-averaging periods.
• Updated discussion of multiple chemical interactions to be consistent with
EPA's recent guidance on chemical mixtures.
• Revised examples using updated risk values from IRIS (1999).
1-13
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1. INTRODUCTION
Section 4:
• Prepared reformatted, streamlined consumption limit tables for each
chemical, using assumptions outlined above (Section 3).
• The definition of "safe fish consumption" was changed from 30 fish meals per
month to 16 fish meals per month.
Section 5:
• Updated chemical-specific information based on IRIS (1999) and other recent
toxicological information on data sources.
• Included additional information on PCBs and dioxin analysis.
Section 6:
• Included new information on georeferencing of fish advisories in the new
Internet version of the National Listing of Fish and Wildlife Advisories
(NLFWA).
Section 7:
• Updated references.
1.6 SOURCES
Information from a wide range of government and academic sources was used in
the development of this document. Current approaches developed by states,
regional groups such as the Great Lakes Sport Fish Advisory Task Force, and
federal agencies including EPA and FDA were reviewed. Section 7 contains a
complete listing of literature sources cited in this document.
In addition, to review the first edition of this document, EPA assembled an Expert
Review Group consisting of officials from several EPA offices, FDA, regional
groups, and the following states: California, Florida, Michigan, Delaware, Illinois,
Minnesota, Missouri, North Dakota, New Jersey, and Wisconsin. A list of the
experts and their affiliations is provided in Appendix A. The Expert Review Group
contributed significant technical information and guidance in the development of
the first edition of this document. Written recommendations made by the experts
were incorporated into the final document. Some members were also consulted
further on specific issues related to their expertise. In a second round of reviews,
this document was circulated to all states, several Native American tribes, and
various federal agencies for comment, and additional modifications were made.
Participation in the review process does not imply concurrence by these
individuals with all concepts and methods described in this document. The Expert
Review Group did not review the current edition of the document because the
1-14
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1. INTRODUCTION
basic risk assessment procedures had already been approved. This third edition
was issued primarily to update new toxicological information for several analytes
and to revise and streamline the consumption limit tables using updated exposure
factors.
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2. RISK ASSESSMENT METHODS
SECTION 2
RISK ASSESSMENT METHODS
2.1 INTRODUCTION
The presentation of risk assessment methods in this section follows the format of
the risk assessment process recommended by EPA for cancer and noncancer
toxicity:
• Hazard identification
• Dose-response assessment
• Exposure assessment
• Risk characterization (U.S. EPA, 1986a,b; IRIS, 1999).
EPA methods follow the outline developed in the National Academy of Sciences
(MAS) report entitled Risk Assessment in the Federal Government: Managing the
Process (MAS, 1983; see Figure 2-1). According to the MAS,
. . . risk assessment can be divided into four major steps: hazard
identification, dose-response assessment, exposure assessment, and risk
characterization. A risk assessment might stop with the first step, hazard
identification, if no adverse effect is found or if an agency elects to take
regulatory action without further analysis, for reasons of policy or statutory
mandate. (MAS, 1983)
Readers may wish to consult the MAS document, Science and Judgement in Risk
Assessment, which updates and expands the 1983 work (MAS, 1994).
Hazard identification is the first step in the risk assessment process. It consists of
a review of biological, chemical, and exposure information bearing on the potential
for an agent to pose a specific hazard (Preuss and Erlich, 1986). Hazard
identification involves gathering and evaluating data on the types of health effects
associated with chemicals of concern under specific exposure conditions (e.g.,
chronic, acute, airborne, or food borne) (U.S. EPA, 1985).
Section 2.2 provides an overview and summary of the hazard identification process
and specific information on hazard identification for chemical contaminants in
noncommercially caught fish. It does not provide detailed guidance on hazard
identification since EPA's Office of Water has already completed the hazard
identification step with respect to fish contaminants. This work was undertaken to
2-1
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2. RISK ASSESSMENT METHODS
RESEARCH
RISK ASSESSMENT
RISK MANAGEMENT
Laboratory and field
observations
Information on
extrapolation methods
Toxicity assessment:
hazard identification
and dose-response
assessment
Research needs identified
from risk assessment
process
Field measurements,
characterization of
populations
Exposure assessment,
emissions characterization
Risk characterization
Development of
regulatory options
Evaluation of public
health, economic, social,
political consequences
of regulatory options
Agency decisions
and actions
Figure 2-1. Elements or risk assessment and risk management
(NAS, 1994).
identify the fish contamination target analytes of concern, as described in Volume
1: Fish Sampling and Analysis (U.S. EPA, 1993a, 1999a) in this guidance series.
This process included an evaluation of information on toxicity, occurrence,
persistence, and other factors. The methods for selecting the highest priority
chemicals as target analytes are described in Volume 1 and summarized briefly in
Section 2.2.1 of this document.
The second step in the risk assessment process is the evaluation of the dose-
response dynamics for chemicals of concern (see Section 2.3). The dose-response
dynamic expresses the relationship between exposure and health effects. To
evaluate this relationship, the results of human and animal studies are reviewed;
the dose-response evaluation may focus on specific types of effects (e.g.,
developmental, carcinogenic) or be designed to encompass all adverse effects that
could occur under any plausible scenario.
The third step in the risk assessment process is exposure assessment (see
Section 2.4). Individual exposure assessments use data on chemical residues in
fish and human consumption patterns to estimate exposure for hypothetical
2-2
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2. RISK ASSESSMENT METHODS
individuals. Population exposure assessments consider the distributions of
exposure in a population. Exposure assessments are then combined with dose-
response data to determine risk.
The final step in risk assessment is risk characterization (see Section 2.5), which
provides an estimate of the overall individual or population risks. Risk
characterization can be used by risk managers to prioritize resource allocation and
identify specific at-risk populations; it is also used to establish regulations or
guidelines and to estimate individual or population risk. In this document, risk
characterization involves developing the risk-based consumption limits provided
in Section 4. When risk characterization is used to estimate individual or population
risk, it provides the risk manager with necessary information concerning the
probable nature and distribution of health risks associated with various co-
ntaminants and contaminant levels.
The importance of describing and, when possible, quantifying the uncertainties and
assumptions inherent in risk assessment has long been recognized, though not
consistently practiced (Habicht, 1992). Uncertainty analysis is particularly critical
in risk characterization and must be performed throughout the risk assessment
process to adequately characterize assumptions in this last step of the process.
Consequently, various sources of uncertainty are described and assumptions are
discussed for each of the four activities that constitute risk assessment.
2.1.1 Other Information Sources
This document focuses on risk assessment as it applies primarily to fish
advisories. EPA has issued several detailed guidelines for conducting specific
portions of the risk assessment process that address the following areas:
• Exposure assessment (U.S. EPA, 1992a)
• Carcinogenicity risk assessment (U.S. EPA, 1986a, 1996b)
• Mutagenicity risk assessment (U.S. EPA, 1986b)
• Developmental toxicity risk assessment (U.S. EPA, 1991a)
• Assessment of female and male reproductive risk (U.S. EPA, 1996a)
• Health risk assessment of chemical mixtures (U.S. EPA, 1986c, 1999a)
• Exposure factors (U.S. EPA, 1990a).
These guidelines were developed by EPA to ensure consistency and quality
among Agency risk assessments. EPA's Risk Assessment Forum is in the process
of developing quantitative guidelines on dose-response assessment of systemic
toxicants. One approach used to estimate reference doses for chronic exposure
toxicity is presented in the Background Documents for IRIS. It is also found in
many EPA publications and has been summarized in papers that discuss risk
assessment within EPA (e.g., Abernathyand Roberts, 1994; Barnes and Dourson,
1988). Relevant sections of each of the above guidelines were consulted in
developing this section, along with other resources cited throughout the section.
Additional references are listed in Section 7.
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2.2 HAZARD IDENTIFICATION
Hazard identification assesses the likelihood that exposure to specific chemicals
under defined exposure conditions will pose a threat to human health. Hazard
identification is often used effectively to determine whether a chemical or groups
of chemicals occurring in a specific exposure situation require action. It has been
narrowly defined for some applications to provide only chemical-specific hazard
data (MAS, 1983). However, in the MAS document, Science and Judgement in Risk
Assessment, the use of an iterative approach to evaluating risk is emphasized,
which entails the use of relatively inexpensive screening techniques to determine
when to proceed to more in-depth evaluations (MAS, 1994). This is analogous, in
practice, to what is already frequently done at the state and local level. The early
stages of risk assessment often include consideration of the existence or likelihood
of exposure to determine the need for further work on a chemical. At the state,
local, and tribal organization levels, administrators and risk managers concurrently
evaluate both the hazard and the occurrence of chemicals to assess whether
sufficient risk exists to justify an investment of time and resources in further action.
Their needs for information to guide further action are, therefore, different from that
of a federal agency, which may evaluate hazards independently of exposure
considerations.
A preliminary risk evaluation typically precedes an in-depth risk assessment
because most states, localities, and tribal organizations do not have the resources
to conduct detailed risk analyses in the absence of information indicating that
health risks may occur. Thus, this section discusses hazard identification as an
approach to making preliminary decisions regarding further action on fish
advisories. This approach is similar to the screening methodology used for the
identification of the 25 target analytes addressed in this guidance series and is
discussed in Volume 1: Sampling and Analysis in this series (U.S. EPA, 2000a).
Although hazard identification is essentially a screening process, it may entail a
complex evaluation of the exposure scenarios and toxicological and biological
properties of contaminants (e.g., bioavailability, degradation, existence of break-
down products and metabolites). Hazard identification ranges in scope from the
use of existing summary data (e.g., IRIS or Agency for Toxic Substance and
Disease Registry [ATSDR] Toxicological Profiles) to a detailed evaluation of each
aspect of exposure and risk; the depth of analysis is usually determined by time
and resource availability. For example, an evaluation of a contaminant's
toxicological properties may include an analysis of all health endpoints likely to
occur in the exposure scenarios of concern. EPA guidance (Habicht, 1992)
describes hazard identification as:
... a qualitative description based on factors such as the kind and quality
of data on humans or laboratory animals, the availability of ancillary
information (e.g., structure-activity analysis, genetic toxicity,
pharmacokinetics) from other studies, and the weight-of-evidence from all
of these data sources.
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Under some circumstances, extensive data collection may be undertaken. For
example, to evaluate carcinogenic risk, EPA has recommended the following
information be reviewed in a hazard identification: physical-chemical properties,
routes and patterns of exposure, structure-activity relationships, metabolic and
pharmacokinetic properties, toxicological effects (including subchronic and chronic
effects, interactions with other chemicals, pathophysiological reactions, and time-
to-response analysis), short-term tests (including mutagenicity and DMA damage
assessment), long-term animal studies, human studies, and weight-of-evidence
(U.S. EPA, 1986a). At the state, local, and tribal organization level, this type of in-
depth analysis is rarely carried out for each health endpoint of a chemical hazard,
due to the time and resources required. Alternatively, databases such as IRIS and
the Hazardous Substances Data Bank (HSDB), which summarize health endpoints
and associated risk values, are inexpensive, readily available, and often consulted
in the development of a hazard profile.
2.2.1 Approach for Fish Contaminants
The hazard identification step in risk assessment of chemically contaminated fish
has been refined by EPA through careful review of the chemical characteristics
considered to be critical in determining human health risk. These parameters are:
• High persistence in the aquatic environment
• High bioaccumulation potential
• Known sources of contaminant in areas of interest
• High potential toxicity to humans
• High concentrations of contaminants in previous samples of fish or shellfish
from areas of interest (U.S. EPA, 1989a).
These characteristics are described in detail in Volume 1: Fish Sampling and
Analysis in this series. Additional information on persistence and bioaccumulation
potential may be obtained from EPA documents such as the Technical Support
Document for Water Quality-Based Toxics Control from the Office of Water (U.S.
EPA, 1991b), which contains a brief description of the bioaccumulation char-
acteristics considered for the development of reference ambient concentrations
(RAC). Readers may also wish to consult the open literature (e.g., Callahan etal.,
1979; Lymanetal., 1982).
2.2.1.1 Toxicological Data—
The toxicity of a chemical to humans can be evaluated based on its acute (short-
term) exposure toxicity and/or chronic (long-term) exposure toxicity. The chronic
toxicity of a chemical is usually of primary concern for environmental toxicants;
however, the varied consumption patterns of fish consumers complicate the
analysis of fish contaminants. This issue is discussed in Section 2.4 in additional
detail. There are a number of databases that contain risk values for various types
of chronic toxicity (e.g., carcinogenicity, liver toxicity, and neurotoxicity). IRIS is a
widely accepted data source because of the extensive review conducted on the
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risk values contained in it. EPA's Health Effects Assessment Summary Tables
(HEAST) are also frequently used (HEAST, 1997). Other relevant databases
include HSDB, the National Cancer Institute's Chemical Carcinogenesis Research
Information System (CCRIS), EPA's GENE-TOX, and the National Institute of
Occupational Safety and Health's (NIOSH's) Registry of Toxic Effects of Chemical
Substances (RTECS). All of the above databases except HEAST are available
through TOXNET.*
2.2.1.2 Contaminant Data-
Information on the prevalence and measured concentrations offish contamination
has been generated through numerous sampling and analysis programs. EPA has
provided a summary of preliminary screening results on the prevalence of selected
bioaccumulative pollutants in fish and shellfish in Volume I of the National Study
of Chemical Residues inFish( U.S. EPA, 1992b). Inaddition, substantial guidance
is provided in Volume 1 of this series on planning a sampling strategy and con-
ducting fish contaminant analyses (U.S. EPA, 2000a).
Likely sources of contaminants are often known to state, regional, and tribal
officials or can be identified through a review of data on manufacturing, toxic
releases, or complaints regarding contamination of food, air, water, or soil.
Recommended sources and lists for obtaining data on probable contaminants
include
• EPA-recommended target analytes (see Table 1-1)
• Chemical releases reported in EPA's Toxics Release Inventory (TRI) database
• The Manufacturers' Index
• EPA priority pollutants
• State inventories of manufacturers and operations
• Chemicals identified in industrial and publicly owned treatment works (POTW)
effluents as nonbiodegradable
• Known spills and contaminants (as reported under the Comprehensive
Environmental Response, Compensation, and Liability Act [CERCLA] to the
Office of Emergency and Remedial Response)
• EPA source inventory for contaminated sediments
• ATSDR's HAZDAT database
• Listing of Superfund (National Priority List) sites
* TOXNET is managed by the U.S. Department of Health and Human Services' National Library of Medicine
(Bethesda, MD). For more information, call (800) 848-8990 (for CompuServe), (800) 336-0437 (for
Telenet), (800) 336-0149 (for TYMNET), or (301) 496-6531 for technical assistance.
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• Common-use chemicals based on practices in the state or region (e.g.,
agriculture or fuels).
This information can be used to describe local waterbodies, incorporating geo-
graphic and source-specific data. The geographic distribution of potential con-
taminants can be used to guide the selection of monitoring sites for sampling and
analysis of potentially contaminated fish.
Volume 11 of the National Study of Chemical Residues in Fish (U.S. EPA, 1992b)
provides an example of how information on the first three characteristics of
chemical contaminants (high persistence in the aquatic environment, high
bioaccumulation potential, and high concentrations of contaminants in previous
samples offish or shellfish from areas of interest) can be summarized to form the
basis for a hazard evaluation. The document summarizes the results of the
National Bioaccumulation Study, correlates contaminant prevalence with sources
of pollutants, and briefly describes the chemical and toxicological properties of 37
chemicals and chemical groups (U.S. EPA, 1992b).
2.2.1.3 Sources of Exposure-
Hazard identification may also include a comprehensive evaluation of all sources
of exposure, including those that augment the primary exposure of concern, to
obtain an estimate of total exposure. For fish contaminants, a comprehensive
exposure evaluation would involve an evaluation of exposures from other sources
such as air, water, soil, the workplace, or other foods, including commercially
caught fish. In some cases, in fact, other routes of exposure may contribute more
to overall contaminant body burden than does contaminated noncommercially
caught fish. It is beyond the scope of this guidance document to provide detailed
direction on evaluating exposures occurring via other media; however, readers are
encouraged to assess other sources of exposures in their hazard evaluations (see
Section 2.4.5.6 for additional information).
If exposure from noncommercially caught fish consumption were added to already
elevated exposure levels arising from other sources, it could produce an overall
exposure associated with adverse health effects. Under such circumstances, a
more stringent fish consumption limit (or some other risk management option) may
be needed. Readers may wish to determine whether such an evaluation is
warranted through consideration of the likelihood that exposures are occurring via
nonfish routes and the availability of data and resources to carry out a
comprehensive exposure evaluation.
EPA's Office of Water, in conjunction with the Interagency Relative Source
Contribution Policy Workgroup, is currently developing guidance on the use of a
Relative Source Contribution (RSC) approach. According to the preliminary
information available on this approach:
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The RSC concept could be used in fish advisory activities. The amount of
exposure from fish consumed is determined along with the estimated
exposure from all other relevant sources (e.g., drinking water, food, air,
and soil) for the chemical of concern. By comparing the overall exposure
with the Reference Dose, it can then be determined whether the amount
of total exposure to the chemical may result in an adverse effect and
warnings can be issued regarding the safety of consuming such fish
(Borum, 1994).
The CERCLA office at EPA, which offers assistance on multimedia assessments
of hazardous waste sites, may also be consulted for information on methods to
estimate background levels of various contaminants. They have developed
guidance documents that may be useful to those readers who plan to conduct
comprehensive exposure assessments.
2.2.2 Assumptions and Uncertainty Analysis
Hazard identification, as described in this guidance, is a screening process used
to select the chemicals and exposure scenarios of greatest concern. As a
screening process, it uses simplifications and assumptions in each step of the
process. Because each aspect of hazard is not examined in its entirety, the
process generates some uncertainty.
Uncertainty is introduced by the variability in persistence and bioaccumulation
potential of chemicals that may occur in untested media. The behavior of
chemicals in all types of media cannot be anticipated. Interactions of the target
analytes in sediments containing multiple chemical contaminants may cause
chemicals to change their forms as well as their bioaccumulation and persistence
characteristics. For example, binding of the target analyte to organic matter may
cause it to become more or less persistent or available for bioaccumulation, or
decomposition may occur, producing metabolites that have significantly different
properties than those of the original target analyte. These chemical and biological
interactions are more likely to occur in a complex system (e.g., a hazardous waste
site), with relatively unstable chemicals, and with metals having multiple valence
states.
The persistence of a chemical in the aquatic environment and its bioaccumulative
potential are based on its physical and biochemical properties. Although the critical
information is available for many chemicals of concern, it is not available for all
chemicals. For example, chemicals that have been recently introduced into the
environment may not be well characterized in terms of their persistence and
bioaccumulation potential. Consequently, there is the potential for under- or
overestimating the risk they pose to human health.
Estimation of chemical toxicity can be a source of significant uncertainty in the
hazard identification process. A toxicity evaluation incorporates data on a variety
of health endpoints and usually requires that human toxicity estimates be derived
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from studies in experimental animals. There are often insufficient data in the
toxicological literature to fully characterize the toxicity of a chemical. Some types
of toxicity are well-described in the toxicological and risk literature. Others, such
as developmental toxicity, neurotoxicity, and immunotoxicity, have only recently
become subjects of intensive research. Although studies of developmental toxicity
date from the 19th century, there has been a dramatic increase in both
epidemiological and toxicological studies in recent years. Consequently, there are
limited data for most chemicals on these types of effects. Uncertainties associated
with toxicity and health risk values (e.g., cancer slope factor ([CSFs] and reference
doses [RfDs]) are discussed in Section 2.3.
The two remaining characteristics of hazard identification (known sources of
contaminants in areas of interest and high concentrations of contaminants in
previous samples of fish or shellfish) are excellent indicators of potential hazard.
A major uncertainty associated with these characteristics arises from the potential
for omitting from sampling programs areas not known to be contaminated. During
an era of limited resources, it is a common, but not necessarily valid, assumption
that known contaminated areas should be the focus of evaluation and action.
Given an array of known contaminated sites, attempts to identify additional
contamination may appear unnecessary. However, it is recommended that readers
conduct a detailed review of potential contamination sources for all waterbodies
before determining whether or not adequate hazard identifications have been
conducted.
Because the goal of the risk assessment process is protection of human health, it
is typically designed to provide the maximum protection against underestimating
risk. Therefore, the hazard identification step in the risk assessment process may
result in the inclusion of chemicals or exposure situations that, later in the process,
are found not to pose significant health risks. This type of approach is taken
because the consequences of underestimating risk, or excluding a chemical that
poses a public health hazard, are potentially more serious than the consequences
of overestimating risk at this early stage of evaluation.
The hazard identification process forms the basis for decisions regarding those
chemicals and exposure scenarios that warrant further analysis. It entails the
collection and evaluation of information regarding toxicity, bioaccumulation
potential, persistence, and prevalence. Although there is uncertainty associated
with this aspect of the assessment, quantitative evaluation of the uncertainty can
best be conducted in later steps in the risk assessment process. Because each
aspect of hazard identification is carried out in more detail in the risk assessment
steps that follow, the uncertainties and assumptions can be better refined and
quantified during subsequent steps. The information generated on toxicity and
exposure in this process also serves as the basis for the subsequent dose-
response evaluation and exposure assessment steps in the risk assessment.
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2.3 DOSE-RESPONSE ASSESSMENT
This section briefly outlines the current EPA methodology for carrying out a dose-
response assessment. Additional information on dose-response evaluations is
available in the references cited in Section 7.
A dose-response relationship expresses the correlation between exposure and
health effects. To evaluate this relationship, the results of human and animal
studies with controlled and quantified exposures are reviewed. This evaluation may
focus on specific types of health effects or be designed to encompass all adverse
effects that could occur under any plausible exposure scenario. Dose-response
evaluations result in the derivation of toxicity values such as cancer potencies and
reference doses.
Actual fish consumption patterns may not correspond well to the typical periods of
exposure studied in toxicity tests (i.e., acute or chronic exposure). Many fish
consumers ingest intermittent doses of varying sizes and may consume fish over
a short period of time (e.g., a vacation) or on a regular basis over a lifetime. The
potentially large, intermittent dose (bolus dose) has not been evaluated in most
toxicity studies. Chronic exposure studies commonly use daily dosing and acute
studies may use one or a few very large doses over a very short time period (e.g.,
2 to 3 days). Short-term dosing is frequently used in developmental toxicity studies
(discussed in Section 2.3.2.3); two of the 25 target analytes have RfDs based on
developmental toxicity (methylmercury and PCBs).
Fish consumption patterns are discussed in more detail in Section 2.4.5.4 and
Appendix B; however, when developing fish advisories, it is important to be aware
that there is no information available on the impact of bolus dosing. The methods
used to calculate fish consumption limits allow the daily RfD to be aggregated over
a period of time (e.g., 1 month) into one or more meals. Thus the consumption
averaged over 1 month corresponds to an average daily dose indicated by the
RfD. However, the actual dose that may be consumed in 1 day can be
approximately 30 times (in the case of a 30-day advisory) the daily RfD.
A bolus dose may not be a problem for many individuals; however, it is a concern
for those who are particularly susceptible to toxicants. For example, a relatively
large single dose may be problematic for those with decreased ability to detoxify
chemicals (e.g., children and the elderly) and those with special susceptibilities
(e.g., persons taking certain medications, children, and pregnant or lactating
women). Potential adverse effects in some groups are noted for many of the target
analytes in Section 5. For example, organochlorines may interact with some
commonly prescribed Pharmaceuticals; consequently, individuals using specific
drugs may find the efficacy altered by large doses of contaminants that interact
with their drug-metabolizing systems. Infants have an immature immune system
and may be less able to detoxify certain chemicals. Children have rapidly
developing organ systems that may be more susceptible to disruption. A MAS
report, Pesticides in the Diets of Infants and Children (MAS, 1993), concluded that
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children up to age 18 are substantially different from adults in the relative
immaturity of their biochemical and physiological functions and structural features.
These differences can alter responses to pesticides, especially during windows of
vulnerability, leading to permanent alteration of the function of organ systems. The
authors, who included pediatricians, toxicologists, epidemiologists, and other health
specialists, concluded that:
Infants and children may exhibit unique susceptibility to the toxic effects
of pesticides because they are undergoing rapid tissue growth and
development, but empirical evidence to support this is mixed
and
Traditional approaches to toxicological risk assessment may not always
adequately protect infants and children (MAS, 1993).
Although the focus of the MAS report was on pesticides (many of the target
analytes are currently or were formerly used as pesticides), much of the analysis
is relevant to other chemical exposures as well. Readers may wish to refer to the
MAS report for a more complete discussion of various related topics of interest
including neurotoxicity in children, various dosimetry scaling methods, and
consumption patterns.
A dose-response evaluation has already been carried out by EPA for the 25 target
analytes addressed in this guidance series. These evaluations resulted in the
calculation of risk values: either CSFs, RfDs, or both. The risk values used in this
work and cited in the toxicological profiles in Section 5 were obtained primarily from
EPA's IRIS database. All data searches were carried out in 1999. For chemicals
lacking IRIS risk values, values were obtained from EPA's Office of Pesticide
Programs (OPP) or EPA's Health Effects Assessment Summary Tables (HEAST,
1997).
A comprehensive dose-response evaluation requires an extensive review of both
the primary literature, including journal articles and proceedings, and the
secondary literature, such as books, government documents, and summary
articles. It is typically very time consuming and requires data evaluation by
toxicologists, epidemiologists, and other health professionals. Because risk values
are available for the target analytes, it is not recommended that readers undertake
further detailed dose-response evaluations for these chemicals. However, new
data are continually being generated that may require evaluation. In addition,
chemicals that are not included in the target analyte list may require analysis. It is
strongly suggested that an evaluation begin with a review of current government
documents on a chemical. In many cases, EPA, FDA, or ATSDR conducts detailed
dose-response evaluations when a chemical is identified as an environmental
pollutant or when new data become available. This may save readers hundreds of
hours of research by providing data and risk values.
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2.3.1 Carcinogenic Effects
EPA has proposed new guidelines for cancer risk assessment (U.S. EPA, 1996b).
These guidelines have not been finalized yet but would supersede the existing
cancer guidelines (U.S. EPA, 1986c). The following discussion presents
information from the existing guidelines that has not changed in the proposed
guidelines and highlights information that has changed. EPA (along with many
other risk assessors) takes a probabilistic approach to estimating carcinogenic
risks. Cancer risk is assumed to be proportional to cumulative exposure and, at low
exposure levels, may be very small or even zero. EPA assumes that carcinogens
do not have "safe" thresholds for exposure; that is, any exposure to a carcinogen
may pose some cancer risk. Carcinogenic risk is usually expressed as a cancer
potency (CSF) value with units of risk per milligram/kilogram/-day exposure. Risk
may also be estimated for specific media. When risks in air and water are provided,
these are referred to as unit risks because they are expressed as risk per one unit
of concentration of the contaminant in air or water.
The cancer slope factor is derived from dose-response data obtained in an
epidemiological study or a chronic animal bioassay. Because relatively high doses
are used in most human epidemiological studies and animal toxicity studies, the
data are usually extrapolated to the low doses expected to be encountered by the
general population. The dose-response data from one or more studies are fit to
standard cancer risk extrapolation models, which usually incorporate an upper-
bound estimate of risk (often the 95 percent upper bound). This provides a margin
of safety to account for uncertainty in extrapolating from high to low doses and
variations in the animal bioassay data (IRIS, 1999). In the existing guidelines, the
model used as a default to calculate the cancer potency is the linearized multistage
(LMS) model. Cancer potency is estimated as the 95 percent upper confidence
limit of the slope of the dose-response curve in the low-dose region. This method
provides an upper estimate of risk; the actual risk may be significantly lower and
may be as low as zero. In the proposed cancer guidelines, straight-line
extrapolation for a linear default is proposed instead of the LMS model. The reason
is that the LMS model gave an appearance of specific knowledge and sophistica-
tion unwarranted for a default model (U.S. EPA, 1996b).
Cancer potencies may be calculated for both oral and inhalation exposure. There
are four major steps in calculating cancer potencies:
• Identify the most appropriate dose-response data
• Modify dose data for interspecies differences
• Develop an equation describing the dose-response relationship
• Calculate an upper confidence bound on the data.
These are described in more detail in the guidelines for cancer risk assessment
(U.S. EPA, 1986a, 1996b) and in texts on risk assessment. Cancer slope factors
are provided for those target analytes that EPA has determined have sufficient
data to warrant development of a value. The values are listed in Table 3-1 and
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discussed in Section 5; they were used to calculate the consumption limits in
Section 4.
As discussed in Section 2.3.2.3, children may have special susceptibilities to some
chemicals and some types of effects. Exposure to a carcinogen early in life may
generate greater risk than exposure later in life. This is due to a variety of factors
including the rapid growth and development ongoing in children and the
proportionally greater consumption by children of some foods. The experimental
literature on this subject is not conclusive and readers may wish to review the MAS
report to obtain additional information (MAS, 1993).
2.3.2 Noncarcinogenic Effects
2.3.2.1 Acute Exposure—
Noncarcinogenic effects that occur over brief periods of time, e.g., a few hours or
days, are considered to be acute exposure effects. They do not necessarily result
in an acute (immediate) response, and so the exposure and response periods must
be considered separately. The pesticide paraquat is an example of a chemical that
usually causes no immediate response to acute exposure but often results in fatal
outcomes after several days or weeks.
Acute exposures have traditionally been considered primarily in the realm of
occupational health or poisoning incidents rather than environmental health
because the brief, low-level exposures associated with most environmental
exposures do not usually result in overt symptoms. The exceptions to this have
been individuals with allergies or chemical sensitivities. However, there has been
a very limited analysis of most environmental pollutants with regard to both the
nature and the critical dose for acute nonlethal effects. Acute exposures are of
concern for fish contaminants due to the ability of fish to bioaccumulate chemical
contaminants to fairly high levels and the relatively large and frequent meals (i.e.,
bolus doses) that may be consumed by sport and subsistence fishers and their
families.
The goal of an acute exposure dose-response evaluation is to identify a threshold
exposure level below which it is safe to assume no adverse health effects will
occur. There are no widely used methods within EPA for setting such exposure
levels. EPA welcomes comments and recommendations on this and other
methodologies.
Mosttoxicological information currently available on acute exposure is in the form
of LD50s from animal studies. These studies identify the (usually single) dose that
was lethal to 50 percent of the study animals via a specific exposure route. The
data are used primarily to give a qualitative sense of the acute toxicity of a
chemical. The information is generally used for purposes of planning industrial and
application processes, transportation, handling, disposal, and responses to
accidental exposures. The data are also used for regulatory purposes and to select
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the less-toxic alternatives among a group of chemical options. LD50s may also be
used to evaluate ecological toxicity.
LD50s are not easily adaptable to an evaluation of the human response to acute
exposures. Because they are focused on the level at which 50 percent of animals
die, they do not provide information on other types of toxic responses, including
those that led to death. Fatal toxic responses may be substantially different from
the responses observed at lower, but still acutely toxic, doses. The LD50 also does
not provide information on the exposure threshold for lethality, which is always
lower (and may be much lower) than the exposure level required to kill 50 percent
of the study subjects. For these reasons, the LD50s have very limited utility in
identifying a threshold for effects of acute exposure. LD50s may, however, provide
comparative information regarding differences in sensitivity between various age
groups or sexes that can be used to evaluate toxicity qualitatively.
Human and veterinary poisoning centers (e.g., Poison Control Centers) are primary
sources of data on acute exposure effects and thresholds. The poisoning data are
limited, however, in many of the same ways in which LD50 data are limited. The
severe responses that often lead to the reporting of an incident do not indicate the
level at which more moderate responses may occur. In addition, the dose is often
not known or is estimated imprecisely. The poisoned individual may have
predisposing medical conditions or may have been exposed concurrently to other
chemicals (including medicines) that affect the nature of the responses.
EPA's Health Advisories also provide some acute exposure information and
guidance regarding 1- and 10-day exposure limits for children with an assumed 10-
kg body weight (available from EPA's Office of Water). Additional information may
be obtained from HSDB. A qualitative summary of acute effects and estimated
human lethal doses is provided for most target analytes in Section 5.
2.3.2.2 Systemic Effects from Chronic Exposure—
Noncarcinogenic effects resulting from multiple exposures occurring over a
significant period of time are also termed chronic exposure effects (IRIS, 1999).
For humans, this usually means exposures over months or years. For animals in
studies used to evaluate human chronic toxicity, the temporal definition of chronic
exposure depends on the species but is usually defined as a significant portion of
the animal's life. Chronic studies are reviewed to determine critical effects for
specific chemicals. The critical effect is the first adverse effect, or its known
precursor, that occurs as the dose rate increases (IRIS, 1999). Subchronic
exposures in toxicity studies (usually 3 months to 1 year) may also be used to
evaluate chronic toxicity.
To protect against chronic toxicity resulting from exposure to contaminants, EPA
has developed RfDs. The RfD is defined as "an estimate (with uncertainty perhaps
spanning an order of magnitude) of a daily exposure to the human population
(including sensitive subgroups) that is likely to be without an appreciable risk of
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deleterious effects during a lifetime" (U.S. EPA, 1987a). The use of IRIS RfDs is
recommended for evaluation of chronic exposure toxicity of the target analytes.
These are listed in Table 3-1 in Section 3 and again in Section 5. Additional chronic
exposure toxicity data for the target analytes are presented in Section 5, with a
brief description of how estimated exposure limits could be calculated based on
chronic toxicity. Note that the RfDs listed in IRIS are subject to change as new
methodologies and toxicological data become available. Readers are advised to
consult the IRIS database to ensure that they are using the most up-to-date toxicity
values.
RfDs calculated for chronic noncarcinogenic effects reflect the assumption that, for
noncarcinogens and nonmutagens, a threshold exists below which exposure does
not cause adverse health effects. This approach is taken for noncarcinogens
because it is assumed that, for these types of effects, there are homeostatic,
compensating, and adaptive mechanisms that must be overcome before a toxic
endpoint is manifested (IRIS, 1999). (Some chemicals such as lead, however,
appear to show nonthreshold noncarcinogenic effects.) It is recommended that
concern be directed to the most sensitive individuals in a population, with the goal
of keeping exposures below calculated RfDs for them (IRIS, 1999). RfDs are
generally expressed in terms of milligrams of contaminant per kilogram consumer
body weight per day (mg/kg-d).
There are two major steps to calculating RfDs: (1) identify the most appropriate
no observed adverse effects level (NOAEL) or lowest observed adverse effects
level (LOAEL) and (2) apply the relevant uncertainty and modifying factors.
1. Identify the Most Appropriate NOAEL or LOAEL
The following hierarchy may be useful in selecting a study from which to use a
NOAEL or LOAEL:
• A human study is preferable to an animal study. When a human study is
unavailable, an animal study is selected that uses a species most relevant to
humans based on the most defensible biological rationale (e.g., pharma-
cokinetic data).
• In the absence of a clearly most relevant species, using the most sensitive
species for the toxic effect of concern is preferable (e.g., exhibiting a toxic
effect at the lowest dose).
• A study with the appropriate exposure route(s) is preferable; oral or gavage is
appropriate for oral exposure.
• A study with sufficient subjects to obtain statistical significance at relatively low
exposure levels is required.
• A recent study identifying adequately sensitive endpoints is preferred (e.g., not
mortality).
• An adequate control population is required.
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• In general, a NOAEL is preferable to a LOAEL. When a NOAEL is unavailable,
the LOAEL that generates the lowest exposure threshold (after the application
of uncertainty and modifying factors) is usually selected.
In addition to the criteria listed, a chronic (lifetime) study is preferable to a
subchronic study (an acute study cannot be used to quantify risks associated with
chronic exposure). It is important that exposure occurs over a significant portion
of the experimental subject's life to parallel a lifetime exposure of the human
population. Issues related to the quality of the study should also be considered in
selecting the most appropriate studies. Additional information on selection criteria
can be reviewed in the IRIS documentation file (U.S. EPA, 1987a).
2. Apply Relevant Uncertainty and Modifying Factors
The calculations for chronic systemic toxicity use the modifying and uncertainty
factors as shown in Table 2-1. In addition, an uncertainty factor may be used when
a chronic study is not available and a subchronic (e.g., 90-d) study is used. This
is generally a tenfold factor (Abernathy and Roberts, 1994; IRIS, 1999). The
product of all uncertainty/modifying factors may range widely depending on the
toxicity database. If a chronic human epidemiologic study is available, the
uncertainty factor may be as small as 1. However, uncertainty factors of 10,000
may be appropriate (Bolger et al., 1990; U.S. EPA, 1990b).
While uncertainty factors address specific concerns, the modifying factor covers
a wider range of circumstances. A common modifying factor adjustment results
from differences in absorption rates between the study species and humans,
differences in tolerance to a chemical, or lack of sensitive endpoint. The default
value for a modifying factor is 1, but may range up to 10 (see Table 2-1).
The uncertainty factor that deals with data gaps has been developed because the
dose-response data often address a limited number of effects and may not
adequately address effects of major concern. (Abernathy and Roberts, 1994). In
some cases there are a number of studies, but the focus of analysis is narrow and
not sufficiently sensitive. In other cases, there is not a sufficient number or breadth
of studies. Other reasons for applying a modifying factor are discussed in the
specific developmental toxicity guidance (U.S. EPA, 1991a); these include data on
pharmacokinetics or other considerations that may alter the level of confidence in
the data. EPA has used the criteria that the following studies be available for a high
level of confidence in an RfD:
... two adequate mammalian chronic toxicity studies in different species, one
adequate mammalian 2-generation reproductive toxicity study, and two adequate
mammalian developmental toxicity studies in different species (Dourson et
al.,1992; U.S. EPA, 1989b).
The uncertainty and modifying factors are divided into the NOAEL or LOAEL to
obtain an estimated dose using the following equation:
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Table 2-1. Uncertainty Factors and Modifying Factors for Estimating
Exposure Limits for Chronic Effects
Uncertainty or
Modifying Factor
General Comments
Uncertainty factor:
human (intraspecies)
Uncertainty factor:
animal to human
(interspecies)
Uncertainty factor: data
gaps
Uncertainty factor:
LOAEL to NOAEL
Modifying factor
Standard
Value
Used to account for the variability of 3 to 10
response in human populations. An
intermediate factor of 3 (1/2 log unit of 10)
may be used if the study examined effects
in a sensitive subpopulation (e.g.,
asthmatics).
Used to account for differences in 3 to 10
responses between animal study species
and humans. An intermediate factor of 3
can be used if appropriate
pharmacokinetic/ dynamic data are
available to justify a reduction in the
uncertainty factor.
Used to account for the inability of any 3 to 10
study to consider all toxic endpoints. The
intermediate factor of 3 (1/2 log unit) is
often used when there is a single data gap
exclusive of chronic data.
Employed when a LOAEL instead of a 3 to 10
NOAEL is used as the basis for calculating
an exposure limit. For "minimal" LOAELs,
an intermediate factor of 3 may be used.
Has been used for differences in 1 to 10
absorption rates, tolerance to a chemical,
or lack of sensitive endpoint. The default
value is 1.
LOAEL = Lowest observed adverse effects level.
NOAEL = No observed adverse effects level.
Source: Adapted from Abernathy and Roberts (1994). Their work also cites: Abernathy et al.
(1993); Barnes and Dourson (1988); IRIS (1999); and Jarabek et al. (1990).
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NOAEL or LOAEL
UF • MF (2~1)
where
RfD = RfD or exposure limit for the target analyte
NOAEL or LOAEL = NOAEL from the selected study
UF = multiplicative product of uncertainty factors
MF = modifying factor.
As a point of reference, EPA has estimated that the RfDs they develop have an
uncertainty spanning approximately 1 order of magnitude (U.S. EPA, 1987a). As
discussed previously, it is necessary to fully characterize the uncertainties and
assumptions that are incorporated in fish consumption limits. A description of the
variability in dose-response results and their impact on fish consumption limits,
descriptions of the data gaps, study limitations, and assumptions are also
important in providing a context for fish consumption limits based on develop-
mental toxicity or other types of toxic effects. It may be useful to review the
description of uncertainties and assumptions associated with dose-response
evaluations provided in Sections 2.3.5 and 5.1.1.12. If this document is the only
source consulted for dose-response data, note that the literature review conducted
for the development of these values was limited to secondary sources such as
ATSDRToxicological Profiles, IRIS, HDSB, and standard toxicological texts (all are
cited in the individual chemical discussions). The list of study characteristics
provided in Section 2.3.2.2 may be useful for identifying data gaps and sources of
uncertainty. The inclusion of this type of information in the risk management
process that follows risk assessment will provide a better overall understanding of
the limitations and uncertainties inherent in the fish consumption limits.
An alternative approach for developing RfDs involves the use of benchmark doses
instead of a NOAEL or a LOAEL. The major limitation of NOAELs and LOAELs is
their subjective reliance on experimental dose spacing and their inability to
adequately account for variability in the dose-response slopes. EPA has developed
guidelines for the use of the benchmark dose approach (U.S. EPA, 1995) and is
in the process of drafting technical guidance for the application of the benchmark
dose approach in cancer and noncancer dose-response assessment.
The benchmark dose approach involves fitting mathematical models to dose-
response data and using the different results to select a benchmark dose that is
associated with a benchmark response, such as a 10 percent decrease in body
weight gain or a 10 percent increase in the incidence of a particular lesion.
2.3.2.3 Developmental Toxicity—
Developmental toxicity has been a recognized medical concern, research subject,
and impetus for restricting exposures of pregnant women to developmental
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contaminants for several decades. However, it is not as well studied as other
health effects such as cancer, and significant gaps in our understanding of
causality and appropriate protective measures remain. Developmental toxicity
incorporates a wide range of effects involving all organ systems in the body.
Prenatal and lactational exposure involves indirect exposure of the developing
fetus; the effective dose may vary with the period of exposure and the specific
chemical. In the past two decades, researchers have determined that the
hypothetical maternal barrier, in the past thought to provide protection for the fetus
during the prenatal period, does not effectively exist. In fact, prenatal exposure
may be especially risky because of the rapid cell replication and differentiation that
occurs in the fetus prior to birth. These same processes also occur at elevated
rates in children and adolescents, causing them to be more susceptible to some
chemical-induced toxicity than adults. Chemical exposures that cause alterations
in the cell replication and developmental processes can lead to serious birth
defects, miscarriages, stillbirths, developmental delays, and a variety of other
adverse effects. A large number of toxic chemicals that have been tested in recent
years have demonstrated developmental toxicity in animal test systems.
Consequently, the exposure of pregnant women to toxic chemicals has become
an area of considerable concern.
Many developmental effects may have environmental causes; however, it is
difficult to establish a causal link in epidemiological studies due to confounders that
arise from the variability in human exposure. It has been estimated that 70 percent
of the developmental defects observed in children are a result of unknown factors
(U.S. EPA, 1991 a); some portion of the 70 percent may be attributable to
environmental exposures.
EPA has studied issues in developmental toxicity and risk assessment for
developmental toxicants over the past two decades and has developed guidance
for evaluating developmental toxicants and establishing health-based exposure
limits. The initial guidance for risk assessment of developmental toxicants was
provided in 1986 (U.S. EPA, 1986b) and has been refined in the current Guidelines
for Developmental Toxicity Risk Assessment (U.S. EPA, 1991a). The
recommended approach uses a NOAEL to calculate an RfD in a manner similar
to that used for the calculation of an RfD based on chronic exposure toxicity. EPA
is also considering use of a benchmark dose approach for developmental toxicants
under some circumstances; consequently, the guidelines may be amended in the
future (U.S. EPA, 1991a). The methodology described in this guidance document
follows the current EPA recommendations. The reader is referred to this and other
sources cited throughout this section and Appendix E for further information on
developmental toxicity risk and risk assessment.
2.3.3 Mutagenicity/Genotoxicity
Mutagenicity and genotoxicity data are not generally used to develop risk estimates
by themselves, although they are frequently used in conjunction with other
information to evaluate other toxicity endpoints (e.g., cancer). There is a wide
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variety of assays designed to assess the mutagenicity of chemicals; however,
there is a limited amount of mutagenicity dose-response data that can be used in
quantitative risk assessment. The majority of data involve in vitro test systems,
which can provide only qualitative evidence of mutagenicity.
The evaluation of weight-of-evidence (WOE) for carcinogenicity, carried out by
EPA for all chemicals having a cancer classification, includes an evaluation of
mutagenicity data. Information on genetic toxicity also needs to be considered
when developing risk values for developmental and reproductive system effects.
Mutagenicity data are summarized in the toxicological profile summaries in
Section 5. Readers are urged to consider this information in reviewing the toxicity
of target analytes. Because information is less readily available on genetic toxicity
and mutagenicity than on other types of risk assessment, and because this type
of toxicity is relevant to evaluating developmental toxicity, a brief summary of the
current EPA guidelines on these types of toxicity has been included in Appendix E.
2.3.4 Multiple Chemical Exposures: Interactive Effects
Most humans are simultaneously exposed to a number of environmental
contaminants. Risk evaluations, however, typically proceed on a chemical-by-
chemical basis. Similarly, the development of risk-based exposure guidelines
typically focuses on the effects of exposure to chemicals individually rather than
as a group. In many cases, the individual exposures and/or risks are then summed
to estimate risks or safe exposure levels for a group of chemicals.
EPA provides guidance on chemical mixtures in risk assessments in Guidelines for
the Health Risk Assessment of Chemical Mixtures (U.S. EPA, 1986c). EPA has
recently published a supplement to the 1986 guidelines (U.S. EPA, 1999a). This
document is intended to reflect the evolutionary scientific development in the area
of chemical mixtures risk assessment. It proposes several different approaches
depending on the nature and quality of the available data, the type of mixture, the
type of assessment being made, the known toxic effects of the mixture or its
components, the toxicologic or structural similarity of a class of mixture or of
mixture components, and the nature of the environmental exposure.
The document proposes that the assessment begins with addressing whether the
type of available data is whole mixture data or mixture component information.
Methods available for whole mixtures are then dependent on whether there is
information directly available on the mixture of concern or only on similar mixtures.
Methods available for component data are dependent on whether there are
interactions data available, whether the components act with a similar mode of
action, or whether the components can be thought of as belonging to a chemical
class (U.S. EPA, 1999a).
A classification scheme is then used to assess the quality and nature of the
available mixtures data. Exposure, health effects, and interactions information is
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then assessed for quality. The results of this assessment determine the type of risk
assessment approach to be used with the mixture. Examples of the approaches
discussed include a mixture RfD or slope factor approach, a qualitative
assessment, a hazard index approach, a weight-of-evidence approach, or
response addition (U.S. EPA, 1999a).
The 1986 guidelines advise the use of the additive approach when data are
available only on individual mixture components. The 1999 guidance also proposes
an additive approach for low exposure levels when interactions information is not
available. For the component chemicals in a mixture that show dissimilar toxicity,
response addition is recommended. Forthe component chemicals in a mixture that
show similar toxicity, dose addition is recommended. Under dose addition, the
general procedure is to scale the doses of the components for potency and add the
doses together; the mixtures response is then estimated for the combined mixtures
dose. Under response addition, the general procedure is to first determine the risks
per the exposure for the individual components; the mixtures risk is then estimated
by adding the individual risks together (U.S. EPA, 1999a).
Section 3 provides a method for calculating exposure limits for multiple chemical
occurrence in single or multiple fish species. The approach is recommended for
use when chemicals have the same health endpoints and mechanisms of action.
The type of information that is often available (acute effects interactions and
mechanisms of action) is not readily applicable to the quantitative assessment of
chronic health risks of multiple chemical exposures (U.S. EPA, 1986c, 1999a).The
guidelines recommend that this type of information be discussed in relation to its
relevance to long-term health risks and interactive effects without making
quantitative alterations in the risk assessment.
The information that may be implied from the toxicological nature of many of the
target analytes is related to the chemical's interaction with basic processes, such
as metabolism. When these functions are altered (e.g., by the induction of
microsomal enzymes), the metabolism of other endogenous or exogenous
chemicals may be altered. This is particularly problematic for individuals using
pharmaceutical drugs to address medical conditions. As the PCB discussion in
Section 5.7 notes, alteration in metabolism of medications may require adjustment
of dosages. This is not a hypothetical problem; exposure to various chemicals has
reportedly resulted in altered response to medications. Information regarding these
types of effects are discussed in Section 5 for the target analytes.
EPA has developed a database to disseminate available information on interactive
effects of chemical mixtures. This database, called MIXTOX, contains summaries
of information from primary studies in the open literature on binary mixtures of
environmental chemicals and pharmaceutical chemicals. Data provided include the
duration of the study, animal species, dose ranges, site, effects, and interactions.
Available MIXTOX information on the target analytes is presented in Section 5. The
majority of data obtained through MIXTOX consisted of the results of acute
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studies. Many studies indicated additive effects. Other types of interactions (e.g.,
inhibition, synergism) were usually not provided. The relevance of this information
to specific waterbodies will depend on the chemical mixtures that are known to
occur, based on fish sampling results. In the absence of quantitative
information on interactive effects, these guidelines suggest the use of an
additive approach to evaluation of chemical mixtures for carcinogens and for
noncarcinogens that are associated with the same adverse health endpoints.
The equation used in this approach is presented and discussed in Section 3.5.
2.3.5 Assumptions and Uncertainties
Numerous assumptions are required to develop risk values from dose-response
data. Uncertainties arise from the assumptions, from the nature of the dose-
response data, and from our imperfect understanding of human and animal
physiology and toxicology. Depending on the quality of the studies, there may also
be uncertainty regarding the nature and magnitude of the effects observed in
toxicological and epidemiological studies. However, evaluation of study quality is
a complex process and involves such diverse topics as animal housing conditions
and pathological evaluations. Often there is not sufficient information provided in
study summaries (either in a journal article or report) to evaluate fully the quality
of the study and the assumption must be made that good laboratory practices and
scientific methods were followed.
Major assumptions that are used in the evaluation of dose-response data are
discussed at length in the EPA risk assessment guidance documents on specific
toxicities (e.g., forcarcinogenicity, numerous assumptions are discussed including
the selection of the dose-response model, use of benign tumors in estimating
response, use of the upper bound estimate of the slope, and use of surface area
instead of body weight to adjust dose [U.S. EPA, 1986a,b,d; 1996b]).
A critical assumption underlying all animal-human extrapolations is that there is a
relationship between toxicity in test animals and the toxicity anticipated in humans.
There can be significant differences in metabolism and other physiological aspects
of study animals and the human population (e.g., absorption, metabolism, and
excretion). Although many of these aspects are well-characterized, the relationship
between interspecies differences and the toxicity of specific chemicals is usually
not known. There is also uncertainty regarding the appropriateness of the test
species for evaluation of a chemical's effects on humans. Generally, the species
of animal that most closely resembles humans in response to the toxicity of a
particular chemical is used in the risk assessment. When such information is not
available (as is often the case), the species of animal that is most sensitive to a
particular effect is used in the evaluation of that effect for a chemical. Although the
existence of a relationship between animal and human toxicity is acknowledged by
most scientists, there is not universal consensus on the nature of the relationship
for many chemicals and endpoints (e.g., male rat kidney toxicity associated with
oc-2-globulin may not be applicable to humans).
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A second critical assumption is the existence of a threshold for most non-
carcinogens and no threshold for carcinogens. The threshold issue is under
evaluation for many chemicals and endpoints (e.g., epigenetic [nongenetic]
carcinogens, developmental effects). Issues of this type will be resolved as more
information becomes available on the basic mechanisms of toxicity and actions of
specific chemicals. Future revisions of this document will provide additional
guidance as it becomes available.
Additional uncertainty regarding dose rate and the duration of exposure is
generated by the use of test animals. Many animal studies are conducted for the
lifetime of the animals; however, the human lifetime is significantly longer than the
2-year study period of the usual experimental subjects (e.g., rats or mice), which
may impact bioaccumulation and toxicity. When human studies are used as the
basis for risk estimates, they are usually of occupationally exposed individuals, who
were exposed intermittently during adulthood over two to three decades rather than
continuously exposed over a lifetime. Often they are not followed into old age,
when many effects become clinically detectable. In addition, human exposures are
often confounded by concurrent exposure to other chemicals. Consequently, the
use of human studies also introduces numerous uncertainties to the toxicity
evaluation process.
Various assumptions are made in most risk assessments regarding the use of
numeric adjustments for extrapolation of study results from animals or human
studies to the general population. The extrapolation models used to estimate
individual or population risks from animal or human studies introduce "margins of
safety" to account for some aspects of uncertainty. These models are designed to
provide an upper bound on cancer risk values and a conservative RfD for
noncarcinogens. Uncertainties arise from the application of uncertainty and
modifying factors in the calculation of RfDs. These factors are based on the best
available scientific information and are designed to provide a safe margin between
observed toxicity and potential toxicity in a sensitive human. The RfD is considered
to be an estimate with uncertainty spanning approximately 1 order of magnitude.
EPA considers the RfD to be a reference point to be used in estimating whether
adverse effects will occur (IRIS, 1999). The IRIS Background Documentation has
provided additional insight into the uncertainty inherent in RfDs:
Usually doses less than the RfD are not likely to be associated with
adverse health risks, and are, therefore, less likely to be of regulatory
concern. As the frequency and/or magnitude of exposures exceeding the
RfD increase, the probability of adverse effects in a human population
increases. However, it should not be categorically concluded that all doses
below the RfD are "acceptable" (or will be risk-free) and that all doses in
excess of the RfD are "unacceptable" (or will result in adverse effects)
(IRIS, 1999).
For carcinogens, the upper 95 percent confidence bound on the linear component
of the linearized multistage model is currently used in estimating a cancer potency
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to introduce a safety margin. It is assumed that this provides a plausible upper
bound estimate of potency in the human population (U.S. EPA, 1986a). EPA'snew
cancer guidelines (which have not been finalized as of this writing) propose using
straight-line extrapolation (U.S. EPA, 1996b).
Many numerical assumptions related to anatomy and physiology are used in
calculating risk values (e.g., average adult body weight of 70 kg, animal dietary
consumption estimates). The application of these assumptions depends on the
type of data being used. These assumptions are typically based on a substantial
amount of information on average or mean values. However, individual variations
within the human population generate uncertainty related to the application of the
assumptions.
Uncertainty is significantly related to the amount and quality of toxicological and
epidemiological data available. There is a greater degree of certainty for chemicals
having human epidemiological studies that encompass a variety of population
subgroups over a dose range. However, this type of data is not usually available.
Uncertainty related to the database is often endpoint-specific. For example, there
may be a substantial amount of data regarding carcinogenic effects but little
information on developmental toxicity. This is the case for many of the chemical
contaminants discussed in Section 5.
Selection criteria for studies are listed for chronic and developmental toxicity in this
section. Where the most appropriate types of data are not available (based on
these selection criteria), there is usually greater uncertainty regarding the risk
values and risk estimates that are calculated. Many of the criteria address the
quality of the studies used to estimate dose-response parameters. Weight-of-
evidence guidelines, also discussed in this section for specific toxicity types,
provide useful insight into the adequacy of the data supporting a risk value.
Bioassays conducted on single cell lines generate greater uncertainty than animal
studies due to their isolation from normal physiological processes. However, some
types of effects can be studied most efficiently using these tests. Various types of
mutagenicity and cellular level assays provide insight into the potential for genetic
damage and damage to specific types of cell systems. These data are very difficult
to interpret in the context of human risk because the relationship between study
results and human effects has not been well-characterized. This type of study is
most often used to support other study results (e.g., positive mutagenicity studies
support animal studies indicating carcinogenicity).
Certain chemicals have such limited data for one or more toxic effects that toxicity
reference values cannot be determined. Some of these chemicals are poorly
characterized for all known/suspected toxicity endpoints. For other chemicals, data
may be well-characterized for certain toxic effects, but inadequate for others. For
instance, the carcinogenicity of organochlorines has been well-characterized in
animals and humans, but other toxic endpoints, including systemic effects and
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reproductive effects, have not been extensively investigated. Limitations for the 25
contaminants in this assessment are described in detail in Section 5.
EPA does not recommend specific factors for modifying toxicity data in cases
where these data are so limited that a dose-response relationship cannot be
determined. However, as the above examples show, lack of toxicity reference
values for a given chemical does not necessarily mean that the chemical causes
no effect. Therefore, readers will need to evaluate if the lack of specific kinds of
toxicity data affect the adequacy of protection afforded by the consumption limit.
For example, if the chemical is a suspected developmental toxicant, but
quantitative developmental toxicity data are lacking, readers may determine that
a consumption limit based on other health endpoints is not sufficiently protective
of women of reproductive age and children.
In summary, uncertainty may be generated by many components of a dose-
response evaluation. Some of these are dealt with quantitatively through the
application of uncertainty factors, modifying factors, or the use of an upper bound
estimate. Others may be referred to qualitatively, through a discussion of data
gaps or inferential information (e.g., studies that appear to show greater
susceptibility at certain ages). The goal of providing the qualitative information on
uncertainty is to give the risk assessor and decision makers sufficient information
on the context and support for risk values and estimates so that they can make
well-informed decisions.
2.4 EXPOSURE ASSESSMENT
This section is meant to provide readers with a brief overview of EPA exposure
assessment methodology. Readers wishing to conduct exposure assessments are
advised to read the more detailed documents listed in Appendix B. Exposure
assessment of contaminants in fish involves six components:
• Chemical occurrences in fish
• Geographic distribution of contaminated fish
• Individual exposure assessment
• Population exposure assessment
• Multiple species exposure
• Multiple chemical exposure.
Each of these components is discussed below.
2.4.1 Chemical Occurrences in Fish
Contaminant concentrations vary among different fish species, size classes within
a fish species, fish tissues, and contaminants present in ecosystems. Chemical
contaminants are not bioaccumulated to the same degree in all fish species. In
addition, chemical contaminants are not distributed uniformly in fish tissues; some
toxicants bind primarily to lipids and others to proteins. Fatty and/or larger fish
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often contain higher organic contaminant concentrations than leaner, smaller fish.
The correlation between increasing size (age) and contaminant tissue
concentration observed for some freshwater fish species (Voiland et al., 1991)
may be less evident in estuarine and marine species (U.S. EPA, 1995; Phillips
and Spies, 1988). Knowing how contaminants differentially concentrate in fish
enables risk managers to advise fish consumers on alternative fishing practices
(consumption of smaller individuals in a contaminated species) and cooking
practices (including skinning, trimming, and cooking procedure) to minimize
exposure.
Volume 1 of this series, Guidance for Assessing Chemical Contamination Data for
Use in Fish Advisories, Volume 1: Fish Sampling and Analysis (U.S. EPA,
2000a), provides comprehensive guidance on cost-effective, scientifically sound
methods for use in fish contaminant monitoring programs designed to protect
public health. It is designed to promote consistency in the data states use to
determine the need for fish consumption advisories. By standardizing protocols
across regions, risk managers can avoid significant differences in advisories when
actual concentrations of chemical contaminants in fish are very similar.
Volume 1 suggests that screening values be compared to annual fish sampling
and analysis data to determine where problems may exist. The document also
discusses sampling design and field procedures for collecting and analyzing fish
and shellfish tissue samples for pollutant contamination. It discusses specific cost-
effective analytical methods, quality assurance/quality control (QA/QC)
procedures, and identifies certified reference materials and federal agencies that
conduct interlaboratory comparison programs. Procedures for data reporting and
analysis that are consistent with the development of the National Fish Tissue Data
Repository (NFTDR) are also included.
Information on contaminant distributions in different types of fish and fish tissues
and across geographic areas is required for a number of reasons. Differential
concentrations of contaminants in fish tissues and across fish species affect fish
consumer exposures due to differences in individual consumption practices. The
geographic origins and modes of transport of chemical contaminants determine the
extent and location of these chemicals in fish. Identifying areas of high
contamination enables readers to choose initial screening sites and focus limited
resources on fisher populations most at risk from consuming contaminated fish.
Many readers will have information on the geographic distribution of contaminants
in fish from their fish sampling and analysis programs. Others may need to identify
areas of likely contamination. This topic is also discussed in Volume 1. This section
briefly reviews likely patterns of chemical distribution based on chemical properties
and other factors. Such geographic information is important in population exposure
assessment and for risk communication; readers are encouraged to develop maps
showing areas offish contamination that, combined with demographic information,
help target exposed fisher populations for additional risk communication and
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outreach efforts. Mapping tools available for tracking locational data on fish
contaminants, fish advisories, or other related data are discussed in Section 6.
2.4.2 Geographic Distribution of Contaminated Fish
The geographic extent of the fish contamination is an important element in
determining the need for further action. These data are also useful in performing
population exposure assessments and risk characterization. Two types of
information are particularly useful: the locations where contaminated fish have
been found and the sources of potential contamination. The first type of information
is provided by fish sampling and analysis programs. When such data are absent,
several available sources can help locate sites of possible contamination by the
target analytes. Section 2.2.1.2 contains a list of sources of information on potential
fish contaminants. Additional information on site selection for fish sampling and
analysis programs is provided in Section 6 of Volume 1.
2.4.3 Individual Exposure Assessment
Individual exposure assessments provide descriptions of the overall, media-
specific, or site-specific exposure of an individual. These may be normative or high
(e.g., highly exposed individual) estimates or be based on actual measurement
data.
Individual exposure assessments use essentially the same equation as that used
with fish contaminants to calculate fish consumption limits, although they solve for
different variables:
C • CR
T7 _ m
m BW
where
(2-2)
Em = individual exposure to chemical contaminant m from ingesting fish
(mg/kg-d)
Cm = concentration of chemical contaminant m in the edible portion of fish
(mg/kg)
CR = mean daily consumption rate of fish (kg/d)
BW = body weight of an individual consumer (kg).
Individual exposure assessments use data on known chemical residues in fish (CJ
and on human consumption patterns (CR/BW) to estimate exposure (EJ for
hypothetical individuals within given populations (see Equation 2-2). Conversely,
the consumption limits described in Section 3 and provided in Section 4 use the
data on known chemical residues in fish (CJ combined with dose-response data
(CSFs and RfDs, which correspond to maximum "safe" exposure) to estimate
maximum safe human consumption rates (CR,im/BW; see Equations 3-1 and 3-3).
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This document uses this equation only to calculate fish consumption limits.
Volume 3 of this series provides additional information on estimating individual and
population exposures for purposes of generating risk estimates used in risk
management decisionmaking. Individual exposure assessment is discussed in this
volume for informational purposes only; it is not used directly in developing the fish
consumption limit tables. Increased detail is provided where information is shared
between individual exposure assessments and consumption limit calculations.
Depending on the geographic region and/or contaminant involved, contaminant
concentrations in fish (CJ are determined by sampling and analysis programs
conducted by public health departments, natural resource agencies, environmental
protection agencies, FDA, EPA, and/or agricultural departments. The consumption
rate (CR) represents the amount of fish an individual in a given population eats in
a day and may be estimated through fish consumption surveys. Finally, the daily
dose is divided by the consumer body weight (BW) to arrive at individual exposure.
By using information on the number of individuals in each exposure category, risk
managers may aggregate exposures determined in individual assessments to
derive population exposure assessments. Population exposure assessments can
allow readers to focus limited resources on those contaminants or areas that may
pose the highest risks to a large number of persons or to particular populations of
interest (e.g., subsistence fishers).
Note: The consumption limits described in this document assume that no other
exposure to any of the 25 target analytes occurs. However, a potentially significant
source of contaminant exposure is the consumption of commercially caught
freshwater, estuarine, and marine fish. Consumption limits for non-commercially
caught fish may not be sufficiently protective of consumers of both commercially
and noncommercially caught fish. It is recommended, therefore, that, whenever
possible, readers take other significant sources of exposure into account when
conducting exposure assessments and/or developing consumption limits.
2.4.3.1 Exposure Variables-
Equation 2-2 uses three parameters to calculate individual exposure (EJ to fish
contaminants from noncommercially caught fish: consumption rate (CR), consumer
body weight (BW), and contaminant concentration (CJ. Equations 3-1, 3-2, and
3-3 in Section 3 also use body weight and contaminant concentration and meal
size (MS) in developing consumption limits. With the exception of Cm, which is
determined through sampling and analysis programs, these parameters are
discussed below.
Body Weight
Both consumption limit and exposure assessment calculations require specific
body weights (usually in kilograms) for individuals in order to derive the con-
taminant daily dose in milligrams contaminant per kilogram consumer body weight
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2. RISK ASSESSMENT METHODS
per day (mg/kg-d). The Exposure Factors Handbook (U.S. EPA, 1990a)
recommends values for average weights for children and adults, based on the
second National Health and Nutrition Examination Survey (NHANESII). Conducted
from February 1976 to February 1980, NHANES II surveyed approximately 28,000
noninstitutionalized U.S. civilians aged 6 months to 74 years. The survey
oversampled population groups thought to be at risk from malnutrition (low-income
individuals, preschool children, and the elderly). Adjusted sampling weights were
then calculated for age, sex, and race categories to reflect body weight values for
the estimated civilian, noninstitutionalized U.S. population. Although EPA
recommends these values for typical Americans, they may not adequately
represent some population groups (e.g., Asian-Americans, who are generally
smaller in stature and have a lower body weight than the average U.S. citizen). If
more accurate data on average body weights of local fisher populations are
available, readers are encouraged to use them in place of the default values.
Table 2-2 lists recommended body weight values for adults, women of reproductive
age (women from 18 to 45 years of age), and children. These values are derived
from data in the Exposure Factors Handbook (U.S. EPA, 1990a); the values listed
for adults are used directly, while the value for women of reproductive age
represents an arithmetic average of three age groups (18-25, 26-35, and 36-45),
and the value for children is an arithmetic average of two groups (children <3 and
children from 3 to <6). A more protective body weight value for women of
reproductive age would be to use the lower 95th percentile body weight of women
ages 18 to 25 years (Blindauer, 1994). In this document, however, a body weight
of 70 kg was used for all adults, including women of reproductive age, to calculate
the consumption limits shown in Section 4.
Meal Size
Meal size is a critical parameter in expressing fish consumption limits, though it is
not used directly in calculating exposure (which is expressed in mg/kg-d).
Consumption limits expressed in terms of meals per given time period are more
Table 2-2. Mean Body Weights of Children and Adults
Mean Body Weight (kg)
Age Group Males and Females
Males Females (Averaged)
Adults
Women of reproductive age
Children <6
78
-
15
65
64
14
70
-
14.5
Source: Adapted from U.S. EPA (1990a).
Bolded values were used in the development of consumption limit tables in Section 4.
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2. RISK ASSESSMENT METHODS
understandable than those expressed in kilograms per day. Meal size estimates
can also be used to calculate peak acute exposures to fish contaminants (although
that information is not used in this document).
Several values for average meal size have been determined through both non-
commercial and commercial fish consumption surveys, although these values may
not be comparable across studies. For instance, some surveys report meal sizes
on the basis of whole, raw fish, while others refer to uncooked fillets. Still others do
not specify whether the value is based on uncooked or raw fish. The average meal
size most often cited is 227 g, or 8 oz (Anderson and Amrhein, 1993; Minnesota
Department of Health, 1992; Missouri Department of Health, 1992; U.S. EPA,
1999a). This meal size corresponds to the value used in the Michigan Anglers
Survey, in which individuals were asked to estimate their average meal size
compared to a picture showing an 8-oz (227-g) fish meal (West et al., 1989). The
same meal size also represents the high-end range used by Dourson and Clark
(1990), which is based on the value used in the EPA Region V Risk Assessment
for Dioxin Contaminants (U.S. EPA, 1988). A discussion of fish consumption
surveys is provided in Appendix B.
EPA suggests using a default value of 8 oz (227 g) of cooked fish fillet per 72-kg
consumer body weight as an average meal size for the general adult population
for use in exposure assessments and fish advisories if population-specific data are
not available. This meal size, however, is not likely to represent higher-end
exposures, where persons consume more than the average amount in a given
meal. These larger meal sizes are important to consider in cases where acute
and/or developmental effects from consumption of contaminated fish are of
concern.
Meal size can also differ for other population groups and must be scaled
accordingly. Children and adolescents, for example, often consume more fish per
kilogram body weight than adults. A national food consumption survey conducted
by the U.S. Department of Agriculture (USDA) was used to scale the adult meal
size value to child meal size values (USDA, 1983). The USDA survey evaluated
consumption patterns of approximately 38,000 U.S. citizens over 3-day periods
from 1977 to 1978 and is the largest consumption survey of its kind that includes
fish. The survey results included meal size data for 10 age groups. Although
respondents included both fishers and nonfishers, relative differences reported
between the age groups were used to approximate differences in average meal
size between different age categories within fisher populations in the current
assessment. For children younger than 4 years old, EPA suggests using a default
meal size of 3 oz (85 g) if population-specific data are not available. For older
children, modifications in consumption limits can be made to tailor limits to their
body weights and consumption patterns. The methodology to do so is discussed
in Section 3.
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2. RISK ASSESSMENT METHODS
Consumption Rate
Although it is necessary to estimate the overall average consumption rate in order
to characterize risk, this information is not necessary to provide risk-based
consumption limits as in Section 4. Consumption rate information is primarily used
to make risk management decisions regarding the allocation of resources and
implementation of various public health protection strategies related to
consumption of contaminated fish. Fish consumption patterns and methods for
evaluating the resulting risks are presented in Appendix B. However, due to the
significant variability in fish consumption among individuals, readers are urged to
conduct their own surveys to determine actual consumption levels when accurate
risk estimates are required.
2.4.3.2 Averaging Periods Versus Exposure Durations—
The exposure duration is the time period over which an individual is exposed to
one or more contaminants. In the case of an individual fisher, the exposure
duration is equivalent to the time interval over which he or she catches and
consumes fish. However, fish consumption is frequently not constant over the time
period of interest for examining certain health endpoints (e.g., lifetime for chronic
effects), particularly for short-term or seasonal recreational fishers. For short-term
or seasonal fishers, periods of consumption must be averaged with periods during
which no consumption occurs to correspond with the time periods over which
chronic health effects are likely to develop. For example, the method usually
employed to obtain a lifetime average daily dose is to divide the cumulative dose
over an individual's lifetime by the number of days in an average lifetime. For
developmental and subchronic effects, the time period over which dose is
averaged is much shorter. Consequently, the time periods of concern chosen for
use in exposure assessments are called averaging periods.
For pollutants with carcinogenic properties, EPA currently assumes that there is
no threshold below which the risk is zero (i.e., for any nonzero exposure, there may
be some increase in cancer risk). There is no current methodology for evaluating
the difference in cancer risks between consuming a large amount of the
carcinogenic contaminant over a short period of time and consuming the same
amount over the course of one's lifetime. EPA's current cancer risk assessment
guidelines recommend prorating exposure over the lifetime of the exposed
individual (U.S. EPA, 1986c) and EPA's proposed cancer guidelines do not
address this issue (U.S. EPA, 1996b). To provide usable and easily understood
consumption guidance, the unit of 1 month was used as the basis for expressing
meal consumption limits for all carcinogenic health endpoint tables shown in
Section 4. The limits for carcinogens are based on the assumption that
consumption over a lifetime, at the monthly rate provided, would yield a lifetime
cancer risk no greater than an acceptable risk of 1 in 100,000.
The likelihood of occurrence of noncarcinogenic effects associated with chronic
exposure is evaluated through the use of RfDs (as discussed in Section 2.3).
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2. RISK ASSESSMENT METHODS
Exposure below the RfD is assumed by EPA to be without appreciable risk over a
lifetime of exposure. Consequently, the relevant averaging time for both
carcinogenic and noncarcinogenic chronic exposure is a lifetime.
As with the carcinogens, the unit of 1 month was used for all tables shown in
Section 4 as the basis for expressing meal consumption limits based on chronic
systemic health effects and developmental effects. The limits for noncarcinogens
are based on the assumption that consumption over a lifetime, at the monthly rate
provided, would not generate a health risk. Although consideration was given to
inclusion of an acute exposure period (e.g., 1 day), insufficient information on 1-
day consumption and acute effects is available to evaluate acute exposure for
many of the fish contaminants at this time.
One or more large meals consumed in a short period (constituting an acute
exposure or "bolus dose") may cause effects substantially different than those
associated with long-term low-level exposures. EPA does not currently have a
methodology that has Agency-wide approval for dealing with high-level short-term
exposures. Consequently, no specific risk values have been provided in this series
to evaluate such exposures (although in future revisions such data may be
available). A qualitative summary of acute toxicity effects of the target analytes is
provided in Section 5. In addition, there are numerous toxicity databases and
books that describe the acute toxicity symptoms of the most common
contaminants. State agencies may refer to these sources or their local poison
control center for guidance on this topic.
Developmental toxicity is often evaluated in animal studies via bolus dose studies,
with exposure over 1 to 3 days, because many adverse developmental effects are
associated with exposures during critical developmental time periods. Severe
developmental effects including stillbirths have been associated with exposures to
high levels of pesticides in foods. Information is provided in a MAS report on
developmental toxicity on special characteristics of infants and children that cause
their exposures and risks to differ from those of adults (MAS, 1993). If very high
exposures are likely to occur, state agency staff are encouraged to consider this
exposure scenario in more detail.
2.4.3.3 Multiple Species Exposures-
Local information on the consumption of multiple fish species and fish
contamination levels can be used to assess exposure and establish consumption
limits for consumers with multiple species diets. Equation 2-2 can be modified, as
follows, to consider consumption of multiple species:
p _ *-^i v mj j ]' ,r\ o\
mj BW
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2. RISK ASSESSMENT METHODS
where
Emj = individual exposure to chemical contaminant m from ingesting fish
species ;'(mg/kg-d)
CmJ = concentration of chemical contaminant m in the edible portion of fish
species ;'(mg/kg)
CRj = consumption rate of fish species; (kg/d)
Pj = proportion of a given fish species in an individual's diet (unitless)
BW = consumer body weight (kg).
Regional or local angler surveys that estimate catch data and measure fish
consumption can provide data on the mix of species eaten by particular popula-
tions. One study, the Columbia River Survey (Honsteadetal., 1971), is described
in Rupp et al. (1979). This survey calculated the total number of each species of
river fish eaten by residents in the area. Although the information is a composite
of fishers and nonfishers, the data could be used to estimate the mix of species
that an average individual in the area would eat. The Columbia River Survey also
includes data on the mix of species consumed by each of 10 individuals who ate
the most fish during the year, which might be used to estimate exposure for high-
risk individuals. Readers may wish to incorporate similar information from local fish
consumption surveys into multiple-species exposure assessments and/or con-
sumption limits.
2.4.3.4 Multiple Chemical Exposures-
Fish can be contaminated with more than one chemical, and individuals can
consume multiple species of fish that contain different contaminants. In these
cases, exposure across species needs to be calculated separately for each
chemical; these exposures can then be combined in a variety of ways to estimate
risks of different health endpoints. Sections 3.4 and 3.5 provide methods for
calculating consumption limits for individuals exposed to multiple contaminants in
a single species and multiple species. Readers also may adapt these calculations
(Equation 2-3) to estimate individual exposure to multiple fish contaminants.
2.4.4 Population Exposure Assessments-
Population exposure assessments are not directly used in developing risk-based
consumption limits. Rather, they are primarily used in risk management (e.g., to
prioritize resource allocation) and to identify particular subpopulations of interest
(e.g., in areas where subsistence fishing is common).
2.4.4.1 Categories of Population Exposure Assessment Information-
Table 2-3 lists the categories of information necessary to evaluate population
exposures. Categories 1 and 2 cover basic demographic data that are often
available from the U.S. Census Bureau. Categories 3 and 4 relate directly to fish
contamination and consumption patterns and should be collected at the local level
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2. RISK ASSESSMENT METHODS
Table 2-3. Categories of Information Necessary for a Population
Exposure Assessment
Category Information
1 Age, sex, and body weight distribution of the population
(demographic data)
2 Average and maximum residence time in an area where exposure
is likely to occur
3 Consumption patterns over the population distribution
4 Levels of contaminants in fish tissue by species, age (size class),
and waterbody
5 General nutritional status of various segments of the population
6 Food preparation and cooking methods
7 Concurrent exposures from other sources to fish contaminants
(e.g., occupational, in drinking water or other foods, airborne, soil)
if possible. Consumption patterns are discussed in greater detail in Appendix B.
Volume 1 of this series provides guidance on sampling and analysis for fish
contaminants as specified in Category 4.
Categories 5, 6, and 7 deal with information, primarily available at the local level,
that is important for overall risk assessment. If local information is absent,
however, data from populations similar to those of concern may be used. If no local
data are available, national data may be used. There are serious limitations to the
use of national data, which are discussed in Appendix B. Using data from other
populations introduces uncertainties. For example, assuming adequate nutritional
status may not be appropriate in an area where nutrition may be impacted
adversely by restrictive advisories. Many chemicals pose greater risks to people
with poor nutritional status (see Section 5 for a chemical-specific discussion).
Consequently, the use of simplifying assumptions may lead to an underestimate
of risk (under other circumstances risks may be overestimated). If poor nutrition is
suspected in populations with high consumption (e.g., sport or subsistence fishers),
obtaining local information is particularly important.
Category 6 deals with information available primarily at the local level on fish
preparation and cooking methods. For some chemical contaminants, skinning and
trimming the fillet as well as cooking can reduce exposure intake. The effect that
fish preparation and various cooking procedures has on reducing contaminant
exposure is detailed in Appendix C.
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2. RISK ASSESSMENT METHODS
Category 7, which deals with multimedia exposure assessment, may be significant
in some areas. Concurrent exposures are important in estimating overall risk and
in determining whether a critical threshold has been reached for threshold effects
(i.e., noncarcinogenic effects). Information should be obtained through local
sampling programs if possible. If local industries contribute to multimedia and
occupational exposures, the overall assessment may be particularly important.
More information on overall exposure assessment and sources of additional
information are provided in Section 2.4.5.6.
This information allows the risk assessor to calculate exposure estimates for a
population. The information may be collected on various groups within the
population (subgroups) who have different consumption rates, culinary patterns,
body weights, susceptibilities, etc.
Identification of susceptible subpopulations is necessary to protect these
individuals adequately. For pregnant and nursing women, women planning to have
children, small children, and people with preexisting health problems, the risk from
consuming contaminated fish may be greater than for healthy men and healthy
nonreproducing women. Some contaminants are particularly damaging during
prenatal or postnatal development. Persons with preexisting health problems may
be particularly susceptible to contaminants that interact with their medications or
that are toxic to the organ systems affected by disease. For these people, low
levels of contaminants may exacerbate their conditions, leading to health effects
not generally experienced by healthy adults. (The special susceptibilities
associated with the various target analytes are discussed in Section 5.) Due to the
above factors, obtaining information on the exposure patterns of susceptible
subgroups is important.
In assembling and reviewing this information, keep in mind the goals of the risk
management activities for the population being evaluated. Decisionmakers should
be aware of the information available and the type of information that will enable
them to identify those at greatest risk. If resources are limited and only one
population subgroup is to be evaluated, evaluating the most highly exposed
subgroups rather than the "average" portion of a population may be advisable. The
highly exposed groups will provide an estimate of the worst-case scenario. These
groups are probably at the greatest health risk (if there is a risk) unless other
groups have more susceptible members. Considering the population exposed at
an "average" level is also important, but, under most circumstances, they will not
be the highest risk group.
Uncertainties and assumptions made in assembling exposure data
should be noted and conveyed to the decisionmakers. It is important to
indicate whether the uncertainties and assumptions are expected to
provide overestimates or underestimates of exposure and risk.
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2. RISK ASSESSMENT METHODS
2.4.4.2 Categorizing Exposure Levels*—
Exposure assessments for a population describe a distribution of individual
exposures. The distribution may be for a geographic area or a particular group of
people (e.g., sport fishers at a particular lake, subsistence fishers in a specific
tribe). It is usually advisable to obtain information on the range of average to high
exposures. Gathering this information allows the decisionmakers to take actions
appropriate for the majority of the population and protective of its most at-risk
individuals. If sufficient resources to evaluate various aspects of exposure exist,
it is recommended that exposure descriptions include the following (Habicht, 1992):
• Individuals at the central tendency and high-end portions of the exposure
distribution
• Highly exposed population subgroups
• General population exposure.
This information can be used to estimate the range of risks from the average risk
(central tendency) to the most at-risk individuals. The 1992 Guidelines for
Exposure Assessment provide detailed and specific guidance regarding
quantification and description for individuals and populations with higher than
average exposure (U.S. EPA, 1992a). This guidance document was the source of
information on the various exposure categories discussed below. As with all
information provided in this document, these recommendations are provided for
reference purposes; state, local, and tribal governments may elect to use any
information they determine is appropriate in establishing fish advisory programs.
Central Tendency
The central tendency represents the "average" exposure in a population. This
value can be derived from either the arithmetic mean or the median exposure level.
Figure 2-2 shows the upper half of a normal population exposure distribution.
When exposure is distributed normally as in the figure, the mean and median will
coincide at the 50th percentile. When the exposure distribution is skewed,
however, the mean and median may differ substantially.
Due to the skewed nature of many exposure distributions, the arithmetic mean may
not be a good indicator of the midpoint of a distribution (e.g., the 50th percentile).
Under these circumstances, a median value (e.g., the geometric mean) may
provide more appropriate information (Habicht, 1992).
* Populations who eat only commercial marine or freshwater fish are not addressed in this guidance
because they are protected through regulation of commercial fish by the U.S. FDA. Exposure values
designed to address consumers of commercially caught fish are not recommended for use in developing
fish advisories.
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2. RISK ASSESSMENT METHODS
Information on the central tendency of a population's exposure may be most useful
in evaluating overall cancer risks and determining the average behavior within a
group. It is not as useful in evaluating noncancer risks because such risks are
based on a threshold for effects. People exposed at levels above the "average"
level may have exposures exceeding the threshold for health effects. If only
"average" levels are considered, the risks to these people will not be considered.
In a normally distributed population, approximately 50 percent of the population will
have exposures above the "average" level.
High-End Portions of the Risk Distribution
The high-end estimates of exposure are those between the 90th and 99.9th
percentiles of the actual (either measured or estimated) distribution. They are
plausible estimates of individual exposures at the upper end of the exposure
distribution. Individuals at the high end of the exposure, dose, and risk distributions
may differ, depending on factors such as bioavailability, absorption, intake rates,
susceptibility, and other variables (U.S. EPA, 1992a). Risks may be reported at
a distribution of high-end percentiles such as the 90th, 95th, and 98th.
Figure 2-2 shows the location of the high-end exposure segment on a normal
distribution. High-end exposure estimates include values falling within the actual
exposure distribution rather than above it. If all factors (e.g., body weight, intake
rates, absorption) are set to values maximizing exposure, an overestimate of
exposure will likely result (U.S. EPA, 1992a). High-end exposure estimates are
very useful in estimating population risks and establishing exposure limits because
they provide a plausible worst-case scenario.
Highly Exposed Subgroups
When a subgroup is expected to have significantly different exposures or doses
from that of the larger population, it is useful to evaluate their exposures.
Bounding Estimates
A bounding estimate of exposure is greater than the highest actual exposure,
corresponding roughly to the upper 99.9th percentile of the population (see Figure
2-2). Bounding estimates are used primarily for screening purposes. Their utility
is in providing the decision-maker with a maximum estimate encompassing the
entire population (Habicht, 1992). They are most useful in eliminating pathways
from further consideration (e.g., if the maximum shows no risk) rather than
determining that a pathway is significant (U.S. EPA, 1992a). Although bounding
estimates are not recommended for use in estimating risks associated with fish
consumption, they may be useful in evaluating the upper bound of risk. Those with
no risk at the upper bound can be eliminated from further concern.
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2. RISK ASSESSMENT METHODS
Typical
Percentile ,
of I 50 /°
Exposure
High End of Exposure
Source: Habicht, 1992.
Figure 2-2. Schematic of exposure categories in upper half of a normal
population distribution.
Data Gaps
The specific information collected for a population exposure assessment will
depend on the goals and resources of the risk managers. Under ideal circum-
stances, detailed local information would be obtained on each category. When
resources are limited, however, assumptions may be necessary for some
categories of information. The EPA publication, Guidelines for Exposure Assess-
ment (U.S. EPA, 1992a), provides the following options for addressing these data
gaps:
• Narrow the scope of the assessment, particularly if the pathway or route with
limited data makes a relatively small contribution to the overall exposure.
• Use conservative assumptions. Conservative assumptions, such as choosing
a value near the high end of the concentration or intake range, tend to
maximize estimates of exposure or dose. If an upper limit rather than a best
estimate is used, express this clearly with the exposure estimate.
• Use models to estimate values and check the conservative nature of
assumptions.
• Use surrogate data in cases where a clear relationship can be determined
between an agent with usable data and the agent of concern.
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2. RISK ASSESSMENT METHODS
• Use professional judgment, especially in cases where experts have years of
observation of similar circumstances.
Data gaps can add significantly to the uncertainty associated with exposure and
risk assessment. Assumptions may be made or data from nonlocal sources may
be used to fill gaps. Selecting health-conservative data will yield health-
conservative exposure and risk estimates; alternatively, selecting less conservative
data will yield less conservative exposure and risk estimates. Decisions concerning
data use will affect risk estimates and may determine where fish advisories are to
be provided.
2.4.5 Uncertainty and Assumptions
Readers must evaluate if the exposure assumptions made in deriving risk-based
consumption limits provide adequate protection to sensitive or highly exposed
populations. Some of the assumptions associated with the exposure parameters
can lead to underestimation of total risk (and therefore overestimation of allowable
consumption). For example, the calculation of exposure to a given chemical may
ignore background sources of that chemical. For chemicals that exhibit health
effects based on a threshold level, the combination of background contaminant
concentration and fish consumption exposure may exceed the threshold. The use
of average fish contaminant concentrations to estimate exposure is another
assumption that could underestimate risk if an individual regularly consumes fish
from a contaminated waterbody.
Exposure assumptions may not always be sufficiently conservative. However,
these assumptions may be balanced by overly conservative assumptions in other
aspects of the assessment. Readers need to judge if the overall margin of safety
afforded by the use of uncertainty factors and conservative assumptions provides
satisfactory protection for fish consumers.
2.4.5.1 Chemical Contaminant Concentrations in Fish-
Exposure quantification requires information concerning fish contamination levels.
Volume 1 contains a discussion of sampling and analysis that provides guidance
on planning and carrying out a sampling program. The document recommends a
two-tiered strategy for monitoring waterbodies for contaminated fish, including:
• Screening waterbodies routinely to identify locations where chemical con-
taminants in fish exceed levels of concern for human health
• Sampling waterbodies intensely where screening has identified elevated levels
to determine the magnitude and geographic extent of the contamination.
Fish contamination varies considerably by waterbody and by fish species and size
class. Therefore, even populations with similar consumption patterns may have
differing exposures, depending on the contaminant levels in the waterbody used
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2. RISK ASSESSMENT METHODS
for fishing. To capture these site-specific distinctions, population exposure
analyses rely on the use of waterbody-specific data from local surveys on fish
contamination. Relevant data from these surveys include levels of contaminants
by fish species and size (length and/or weight).
Accurate determination of the chemical concentrations in fish is an important area
of uncertainty that is discussed in detail in Section 8 of Volume 1 in this series. The
limit of detection (LOD) for each of the 25 target analytes is given in the footnotes
of the consumption limit tables in Section 4 and in Appendix F.
2.4.5.2 Dose Modifications Due to Food Preparation and Cooking—
EPA recommends the use of dose modification factors for setting
health-based intake limits only when data on local methods of prepara-
tion and their impact on contaminant concentrations are available.
Several sources of uncertainty are associated with the dose modification factors
presented in this guidance. Preparation methods are frequently unknown. The
effectiveness of different preparation and cooking techniques in reducing con-
taminant concentrations varies greatly. In addition, information is limited regarding
the toxicity of the degradation products generated during the heating of con-
taminated fish. Percentage reductions observed at one level of contamination may
or may not be expected to hold true for different levels of contamination. These
sources of uncertainty could lead to either under- or overestimates of exposure.
Additional discussion on dose modification is provided in Appendix C.
2.4.5.3 Body Weight-
The estimates for body weight use several assumptions that affect the accuracy
of the exposure assessment. First, the figures for body weight are taken from data
collected in the late 1970s. Body weights can vary dramatically over time, and
therefore the values may be an over- or underestimate of current body weights. In
addition, average body weights were not distinguished for various ethnic
populations. For example, Southeast Asian-American subsistence fishers may
have slighter body frames and lower body weight than the general U.S. adult
population. Compared to other assumptions, however, body weight values are
associated with relatively low variability and uncertainty.
2.4.5.4 Consumption Rate and Averaging Period-
Fish consumption data are a necessary component of a population exposure
assessment. Ideally, fish consumption information will include descriptive
demographic information on the size and location of the fishing population using
specific waterbodies; the age and sex of those consuming the fish; the size and
frequency of the meals (over the short and long term); and the species of fish
caught, portions of the fish consumed, and methods of fish preparation and
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cooking. This section discusses the selection of fish consumption data and
presents results obtained in numerous studies.
In general, fish consumption studies describe:
• Species of fish consumed by various subgroups within a population
• Temporal patterns of consumption
• Variety of preparation and cooking methods used by different populations.
Many studies provide some, but not all, of the above data.
Consumption patterns may differ significantly both within and between populations.
Studies of fish consumption indicate that some groups within the general U.S.
population may consume considerably greater quantities of fish than other
members of the population.
This document focuses on noncommercial fishers (i.e., people who fish and
consume their catch) and the people with whom they share their catch. This sub-
population may include sport fishers and subsistence fishers. Sport fishers include
all noncommercial fishers who are not subsistence fishers. (They have also been
referred to as recreational fishers.) Subsistence fishers, as previously defined,
include people who rely on noncommercial fish as a major source of protein.
Subsistence fishers may also catch fish for commercial sale; however, this activity
comes under the jurisdiction of the FDA and is not considered in this document.
There is often not a clear distinction between sport and subsistence fishers. Many
individuals would not consider themselves subsistence fishers but do rely on non-
commercially caught fish for a substantial portion of their diet. The mean or median
estimates of consumption rates and patterns generally address the more casual
sport fisher; the high-end estimates (upper percentiles) and patterns address the
consumers at greater risk. In many of the older surveys, the high-end estimates
were used as estimates of the consumption rates for all subsistence fishers. These
estimates, however, may be inaccurate because some surveys excluded subpopu-
lations that tended not to register for fishing licenses.
The two most sensitive variables involved in calculating individual exposure often
are consumption rate and averaging period. Consumers of noncommercially
caught fish differ immensely in their consumption habits. Some may consume fish
for 1 week during a year or for several weekends each year (e.g., as recreational
or sport fishers). Others may consume fish for much longer periods during a year
(seasonal fishers) or may rely on fish year-round as a major part of their diet
(subsistence fishers). Within these groups, some individuals are more susceptible
to contaminants, including women of reproductive age, children, and persons with
preexisting health problems.
Short-term recreational and seasonal fishers are assumed to be exposed to
contaminated fish for only part of the year. Recreational vacation fishers are those
who eat fish only a short time during the year. Seasonal fishers are often those
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who live near a lake or river, who fish regularly throughout a season (e.g., summer
fishing, winter ice fishing), and who eat their catch throughout the season but do
not rely on fish as a major dietary staple during the rest of the year. Sport fishers
have been shown to have higher fish consumption rates than the general U.S.
population (U.S. EPA, 1989a); the potential for large exposures over short time
periods makes them especially susceptible to acute, developmental, and
subchronic health risks as compared to nonfishers.
Subsistence fishers eat fish as a major staple in their diets for a greater percentage
of the year than do recreational fishers. In addition, subsistence fishers may
prepare fish differently than do other groups; they may use the whole fish in soups
or consume more highly contaminated tissues, such as the liver, brains, and
subcutaneous fat. Both their longer exposure durations and consumption habits
make many subsistence fishers more likely to be affected by cancer and adverse
chronic systemic, developmental, and reproductive health effects resulting from
fish contaminant exposure than those who do not fish or fish for shorter periods of
time. Some populations that may subsist on noncommercially caught fish year-
round, including certain Native American tribes, may be at higher risk (see Section
1.3). In addition, certain recent immigrants accustomed to self-sufficiency and
fishing (particularly Asian-Americans) and economically disadvantaged populations
may be at risk since much of their fishing might be expected to occur in more
urbanized areas with higher levels of water pollution.
Any estimates of typical fish consumption patterns in a population include certain
assumptions. West et al. (1989) described variations in fish consumption in
communities in Michigan by ethnicity, income, and length of residence. In general,
African Americans and Native Americans ate more fish than Caucasians; older
individuals ate more fish than younger individuals; individuals with lower incomes
tended to consume greater quantities of fish than individuals with higher incomes;
and longer-term residents of the communities tended to consume more fish than
other individuals. To the extent that members of the target population have
characteristics associated with higher-than-average consumption, the
recommended consumption values may underestimate their consumption. Unless
surveyed specifically, subsistence fishers may be underrepresented by available
surveys. Surveys associated with the issuance of fishing licenses are traditional
mechanisms used in surveying fish consumption behavior; however, subsistence
fishers may not apply for fishing permits or licenses. For example, Native
Americans on reservations do not need fishing permits, and often times other
groups (e.g., recent immigrants or the elderly) may not know that they need to
have a license or find them too expensive to buy.
In addition, fish consumption limits that are based on single species for single
chemicals do not account for exposures from multiple chemicals contaminating a
single species or for multiple species diets. Consumption limits that focus on a
single waterbody do not account for the possibility that consumption can occur
from a variety of waterbodies. Single-species consumption limits also do not
address related species that may be contaminated but were not sampled. Such
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2. RISK ASSESSMENT METHODS
consumption limits could seriously underprotect persons who eat a variety of fish
species from a number of waterbodies. Readers need to decide if consumption
limits have a wide enough margin of safety to protect such consumers.
Other methodological assumptions may also lead to increased uncertainty. The
calculation of consumption limits that express allowable dose as a number of
meals over a given time period may neglect potential acute effects if consumption
occurs over a very short time period. For example, a meal limit of two meals per
month conceivably could be interpreted by consumers to mean that two meals on
1 day in a given month is allowable; this behavior could lead to short-term acute
effects. This could be avoided by always expressing the consumption in terms of
the time interval in which one meal may be consumed, (e.g., one meal per 2 weeks
rather than two meals per month).
The use of averaging periods treats large, short-term doses as toxicologically
equivalent to smaller, long-term exposures when comparing exposure to the
toxicity reference value. This assumption may underestimate the potential toxicity
to humans if the toxicity depends on a mechanism sensitive to large, intermittent
doses. (This may occur more often with acute and developmental effects than with
other effects.)
The averaging period of 1 month used in this document is based primarily on the
types of health data currently available and the risk assessment methods
recommended by EPA.
2.4.5.5 Multiple Species and Multiple Contaminants—
As discussed above, individuals often eat more than one species of noncommer-
cially caught fish in their diet. If consumption limits or exposure assessments
consider only a single-species diet, exposure from contaminated fish could be
underestimated if other species have higher concentrations than the species under
consideration. On the other hand, an exposure assessment may be overprotective
if an individual's diet is a mix between contaminated and uncontaminated species.
Use of local information to the extent possible to characterize mixed diets can
prevent some of this uncertainty.
An individual may consume a given species that is contaminated with multiple
chemicals, or may consume several species, each with different contaminants, or
both. In these circumstances, exposure assessments that examine contaminants
individually in individual species will underestimate exposure. This situation may
be avoided by using Equation 2-3 in Section 2.4.3.3 for multiple species exposures
and characterizing exposure to all known contaminants for a given individual.
These exposure values can be used in methods described in Sections 3.4 and 3.5
to set consumption limits based on multiple species and multiple contaminants.
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2. RISK ASSESSMENT METHODS
2.4.5.6 Other Sources of Exposure—
The methods described in this guidance consider exposure primarily from
consumption of noncommercially caught fish. This approach may lead to an
underestimation of exposure and, consequently, an underestimation of risk for
some contaminants. Additional background exposure may cause individuals
exposed to fish contaminants through other contaminant sources (e.g., other foods
including commercially caught fish, drinking water, inhalation, or dermal contact)
to experience adverse health effects and/or increased cancer incidence, even if
they abide by the consumption rates recommended in fish consumption advisories.
State agencies are encouraged to use available information on other sources of
exposure whenever possible in setting consumption limits or to set the limits so
that the allowable consumption accounts for only a fraction of the total allowable
daily dose. These approaches would allow a margin of safety to guard against the
potential for background exposure leading to exceeding the contaminant
thresholds and/or maximum acceptable risk levels.
Nonfish Sources of Exposure
People may be exposed to one or more of the target analytes through sources or
pathways other than noncommercially caught fish. These pathways include
contaminants found in or on commercially caught fish, other food, drinking water,
air, or other materials (e.g., soil or sediment).
Contact may often occur via more than one route of exposure (e.g., ingestion and
dermal contact with contaminants in soil). The possibility of exposure via other
pathways dictates that caution be used in setting health safety standards that do
not take these other sources into account. The total exposures may cause the
individual to exceed a safe exposure level, even though the exposure via fish
consumption alone may be safe.
EPA is currently developing a relative source contribution method, which can be
used to evaluate the amount of exposure contributed from various sources. The
RSC method can be used to compare total contaminant exposure to that
contributed by a specific source (e.g., fish); it is also useful in evaluating the total
exposure from all sources. Information on the relative contribution offish to overall
exposure can be used to develop advisories that recommend sufficiently low
exposure to ensure that total daily exposure is below an established targeted
exposure level (e.g., an RfD). It is anticipated that information regarding the RSC
method will be incorporated into future revisions of this document.
If state agencies have information about other pathways that may contribute
significantly to exposure, then risk assessors are encouraged to use this
information to calculate an appropriate total exposure limit. An alternative approach
may be appropriate when nonfish exposures are suspected but have not been
quantified. Depending on the magnitude of the suspected nonfish exposure, the
fish advisory intake limits may be set at a level that accounts for some fraction of
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2. RISK ASSESSMENT METHODS
the total allowable daily dose (e.g., 10, 20, or 30 percent). This allocates to the
nonfish exposures the remaining percentage of the total exposure limit. The goal
of both of these strategies is to ensure that the total pollutant exposure does not
exceed the predetermined exposure limit.
One state program raised concerns that this series focuses on reductions in
exposure via fish when exposures via multiple media may be occurring. However,
it is important to note that, although exposure reductions can theoretically be made
in any contaminated media, fish consumption may be the only source that can be
readily reduced. It may not be possible to reduce air, drinking water, or other
contaminant levels quickly, yet fish advisories have the potential for rapid exposure
reduction in a population. Because fish consumption may contribute significantly
to overall exposure for some population groups, modified consumption patterns
may reduce overall exposure considerably. The relationship between fish and other
contaminant source contributions to overall exposure should be communicated to
risk managers so that both short- and long-range planning for exposure reduction
can occur.
Estimating Total Exposure
The following discussion of exposure calculations is similar to that provided in
Section 2.4.3 for individual exposure assessment. Exposure assessments provide
descriptions of the overall, contaminant-specific, media-specific, or population-
specific exposure of an individual or similarly exposed group. The following
equation may be used to express exposure in a manner (mg/kg-d) that can be
easily compared to an RfD or used to calculate cancer risks:
C • CR
F
r>w A
r>W
where
ET = exposure from all sources (mg/kg-d) to contaminant (m)
Cm = concentration in the edible portion of fish (mg/g)
CR = mean daily consumption rate of fish (g/d)
BW = average body weight of the group (kg)
EA = exposure from air sources (mg/kg-d)
Ew = exposure from water sources (mg/kg-d)
EF = exposure from nonfish food sources (mg/kg-d)
E0 = exposure from other sources such as soil (mg/kg-d).
The equation expressing average daily consumption per kilogram in Appendix D
can also be used to express fish-borne exposure (the Cm, CR, and BW portion of
the equation). If the concentration in fish tissues is reduced due to preparation or
cooking, the Cm value should be modified accordingly. Note that loss of
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2. RISK ASSESSMENT METHODS
contaminants, with a proportional loss of fillet weight, will not change the
concentration, which is expressed in milligrams of contaminant per kilogram offish
(mg/kg). Finally, the daily exposure (mg/d) is divided by consumer body weight
(BW) to arrive at individual daily intake (mg/kg-d).
Body weights for various age groups of consumers are summarized in Table 3-5.
If high estimates of body weight are used (e.g., adult male values), the risks and
fish advisories will be less health conservative. If lower body weights are used
(e.g., for small women), the risks and fish advisories will be more health conserva-
tive. When children's exposure is evaluated separately, their body weights should
be used in conjunction with their estimated consumption rates. Risk managers may
wish to consider the population they seek to protect with their fish advisories and
whether they wish to protect the most at-risk groups in selecting a body weight.
The selection of a body weight value will not have a substantial impact on the final
values because the differences in body weight are relatively small (less than a
factor of 2) compared to the uncertainties associated with most toxicological data.
Methods for estimating exposure to multiple contaminants and multiple fish species
are discussed in Section 3 and equations are provided. These equations for
individual exposure estimates can also be used for populations with similar
exposure characteristics.
The type of exposure information collected and evaluated will depend on the
resources and goals of the fish advisory program. Under ideal circumstances,
pollutant levels would be evaluated in all media to which individuals may be
exposed. For example, drinking water contaminant levels may be evaluated by the
local water purveyor on a regular basis, and this information can be used to
estimate waterborne exposure. When pesticides are the subject of concern, the
evaluation may be more difficult because the levels present in food are not
evaluated frequently at the local level. In addition to providing necessary
information for the development of fish advisories, a total exposure assessment
may highlight nonfish sources of exposure that merit attention.
Summarizing Exposure Information
Table 2-4 is a template for use in summarizing exposure information. It contains
entry areas for fish exposure and exposure via other media. Risk assessors and
managers may wish to use this template to organize their exposure data for
various population groups or subgroups by chemical. The table is designed to
organize data obtained from a specific location (e.g., an area adjacent to part of a
waterbody or surrounding an entire waterbody). It is anticipated that the information
entered in this table would be organized according to population subgroups with
similar risk characteristics (i.e., a separate table should be pre-pared for children,
women, etc).
As noted earlier, exposure levels may differ among subgroups within the fish-
consuming population, depending on the species offish that are caught, the
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Table 2-4. Exposure Data Template
Location:
Population Subgroup (e.g., children, women 18-45 yr):
Population Size:
Body Weight:
Contaminant
(level)
Fish Exposure
Estimates
(mg/kg-d)
Central
High
Enda
Other Exposures
Air (mg/kg-d)
Central
High
End
Water (mg/kg-d)
Central
High
End
Food (mg/kg-d)
Central
High
End
Other (e.g., soil)
(mg/kg-d)
Central
High
End
Subtotal of
Other Exposures
(mg/kg-d)
Central
High
End
Total of All
Exposures
(mg/kg-d)
Central
High
End
10
33
CO
7;
m
m
m
O
o
aRisk assessors may wish to use a bounding estimate rather than a high end estimate (or both).
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2. RISK ASSESSMENT METHODS
quantity of fish consumed, and the method of preparation and cooking used. In
some cases, other factors will also affect exposure (e.g., seasonal changes in
contaminant levels, the age of the fish). For purposes of risk assessment,
specifically targeted risk information is obtained when the exposure of a population
group is the same and their susceptibilities to the chemicals of interest are the
same.
Estimates may be made for average, high-end, or upper-bound exposures within
a population group. The use of average exposure values is not recommended
because approximately one-half of the population will have exposures greater than
the average (by definition). High-end estimates maximize the protection of public
health. Upper-bound values may yield unrealistically high estimates of exposure
and risk and are more appropriate for screening purposes than for risk
assessment. Depending on the characteristics and needs of the fisher population,
risk managers may elect to use the values they deem most appropriate.
The template provides entry areas for central tendency, high-end exposure, and
bounding estimates. By including these categories of information, risk assessors
can calculate a wider range of risk estimates and risk managers will have more
complete information on which to base decisions concerning appropriate fish
advisories. It may not be practical, however, to do three levels of calculations for
each area, group, and contaminant. Table 2-4 does not contain a separate entry
column for dose modifications due to cooking or cleaning. If these activities are
known to reduce exposure, risk assessors may enter appropriately reduced
exposure values to account for the dose reduction (see Appendix C for additional
information).
The information entered in Table 2-4 will be used with risk values to calculate risks.
For this reason, body weight, an essential component of risk calculations, is
included. It is assumed that body weights corresponding to the population of
interest will be used. For example, if specific calculations are to be carried out for
women exposed to mercury, then a separate exposure table (or entry) for women,
using appropriate consumption and body weight values, is advisable. Similarly, if
risks are to be estimated for children or separate advisories developed for this
group, information concerning children's exposure would be entered separately.
Exposures to contaminants from media other than fish may vary considerably for
children in comparison to adults. Children have higher intakes of food, drinking
water, soil, and air in relation to their body weight than do adults (MAS, 1993). In
particular, infants consume significantly greater amounts of fluid than older children
and adults. If contaminants are known or thought to occur in water supplies, infants
may be a subpopulation for whom a separate analysis would be warranted,
especially if water is used to mix formula. If the contaminant of concern is
concentrated in human breast milk, breast-fed infants may be at greater risk.
Any exposure information that will modify the total exposure of the target
population may be entered in the template to indicate differences from the larger
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2. RISK ASSESSMENT METHODS
population. Situations such as workplace exposure, high periodic fish
consumption, or other occurrences can be noted and evaluated for their impact on
overall health and risk.
2.5 RISK CHARACTERIZATION
In general, the risk characterization step of the risk assessment process combines
the information for hazard identification, dose-response assessment, and exposure
assessment in a comprehensive way that allows the evaluation of the nature and
extent of risk (Barnes and Dourson, 1988). Risk characterization can be used by
risk managers to prioritize resource allocation and identify specific at-risk
populations; it is also used to establish regulations or guidelines and to estimate
individual or population risk. In this document, risk characterization has been used
to develop the risk-based consumption limits provided in Section 4. The methods
involved in developing consumption limits are described in detail in Section 3 and
are not repeated here. When risk characterization is used to estimate individual or
population risk, it serves to provide the risk manager with necessary information
concerning the probable nature and distribution of health risks associated with
various contaminants and contaminant levels.
Risk characterization in general has two components: presentation of numerical
risk estimates, and presentation of the framework in which risk managers can
judge estimates of risk (U.S. EPA, 1986a). A characterization of risk, therefore,
needs to include not only numerical characterizations of risk, but also a discussion
of strengths and weaknesses of hazard identification, dose-response assessment,
and exposure and risk estimates; major assumptions and judgments should be
made explicit and uncertainties elucidated (U.S. EPA, 1986a).
Numerical presentations of risk can include either estimates of individual risk or
risks across a population. For example, for cancer risks, numerical estimates can
be expressed as the additional lifetime risk of cancer for an individual or the
additional number of cases that could occur over the exposed population during
a given time period. Numerical risk estimates can also be expressed as the dose
corresponding to a given level of concern (U.S. EPA, 1986a). These values can be
used to estimate the environmental concentration or contact rate below which
unacceptable health risks are not expected to occur. For the determination offish
advisories, the environmental concentration takes the form of screening values
(i.e..contaminant concentrations in fish, as discussed in Volume 1) and the contact
rate takes the form of risk-based consumption limits for specified populations.
Additional factors to be considered in risk characterization include:
• Possible exposure to the fish contaminant(s) from additional sources (e.g., air,
water, soil, food other than fish, occupational activities)
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2. RISK ASSESSMENT METHODS
• Characteristics of the population that may cause them to be more susceptible
than the general population due to exposures to other toxicants, their general
health and nutritional status, or their age
• An absence of sensitive study data for significant health endpoints such as
developmental abnormalities, neurotoxicity, and immunotoxicity
• Recent toxicological study results indicating potential health risks not
considered in the current risk values
• Information from local medical practitioners indicating likely risk-related health
effects
• Economic, nutrition, or other hardships that may result from fishing restrictions.
Most of the factors listed above may lead a state agency to select more health-
conservative risk values. For example, when information concerning a population
(or subgroup) indicates that they have poor nutritional status that may increase
their susceptibility to a local contaminant, state agencies may elect to modify the
risk values they are using directly to provide an additional "margin of safety."
Although the RfDs are designed to protect the most sensitive individuals, state
agencies have discretion in determining the appropriate approach to protecting the
public health of the people they serve.
The last factor listed above is an important risk management consideration. Use
of health-conservative risk values will result in more restrictive fish advisories,
which may have serious impacts on local populations.
In many cases the advantages and disadvantages of selecting specific risk values
will affect members of communities in different ways. Groups at highest risk will be
the most likely to gain from being alerted to health hazards (if they choose to take
protective action). Alternatively, groups with relatively low risks may unnecessarily
avoid consumption of food or participation in the sport of fishing, even though these
may have overall benefits to them (i.e., the risks may be outweighed by the
benefits).
There will invariably be tradeoffs between protection of public health and unwanted
impacts of consumption restrictions. In some cases, the benefits of advisories may
be a generally agreed-upon community value (e.g., preventing relatively high risks
to pregnant women). Other cases may be less clear, especially when the scientific
evidence on risks is limited. Decisionmakers are urged to consider the scientific
information, fish consumption patterns, community characteristics, and other local
factors carefully, along with potential positive and negative impacts of their
decisions, when selecting risk values for screening or establishing advisory limits.
Involving the affected communities in the decision-making process may be
advisable under most circumstances.
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2. RISK ASSESSMENT METHODS
See Appendix D for EPA's guidance for risk characterization, which discusses the
basic principles of risk characterization.
2.5.1 Carcinogenic Toxicity
In this guidance series, screening values are defined as the concentrations of
target analytes in fish tissue that are of potential public health concern and that are
used as standards against which levels of contamination can be compared. For
carcinogens, EPA recommends basing screening values on chemical-specific
cancer slope factors. Screening values are used to establish the concentration in
fish that can trigger further investigation and/or consideration of fish advisories for
the waterbodies and species where such concentrations occur. The method for
calculating screening values is given in Volume 1 of this series.
2.5.1.1 Individual Risk-
Using cancer slope factor and exposure data in mg/kg-d, cancer risks are
calculated using the equation:
Lifetime risk = exposure x cancer potency (2-5)
where
exposure = total exposure to a single contaminant from all sources
(mg/kg-d)
cancer potency = upper bound of the lifetime cancer risk per mg/kg-d.
Note that cancer risk can be estimated for individual sources of exposure. Use of
the total exposure value yields an estimate of lifetime cancer risk from all sources
of a single contaminant. The resulting value is the upper bound of the estimated
lifetime cancer risk for an individual or for a group with the same exposure level.
Different exposure levels may be used in the above equation to calculate risks for
different groups within a population having differing consumption rates, body
weights, etc.
EPA cancer slope factors are based on an assumed exposure over a lifetime;
consequently, adjustment for differences in consumption and body weight in
childhood may not be necessary. Based on the occurrence of some childhood
cancers, it is suspected that exposure to some chemicals may not require a lifetime
to generate risk. However, carcinogenic toxicity tests in animals are usually
conducted for the lifetime of the animal. Consequently, it is not possible to
determine, for most contaminants, if there are risks that may be generated with a
brief exposure duration. This remains an area of uncertainty. When human data
are available, which is relatively rare, impacts on children are often better
understood (e.g., risks are well known for ionizing y radiation). In addition, it is
worth noting that the lifetime cancer risk equation is the linear approximation that
is reasonable for low doses/risks, but that cancer risk cannot exceed 1 and as it
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2. RISK ASSESSMENT METHODS
approaches 10~2, the exponential form of the equation is needed to make accurate
estimates.
2.5.1.2 Population Risk—
The estimated population cancer risk is calculated by multiplying the number of
people in an exposure group (with the same exposure) by the lifetime cancer risks
calculated from the equation above. The population risk equation is:
(population cancer risk) = lifetime risk x (size of exposed population). (2-6)
For example, if 5,000 people are exposed at a risk level of one per thousand (1 x
103) (per lifetime), the overall risk to that population is five additional cancer cases
(5,000 x 1 x 10~3 = 5) over the background level.
Because risks always vary across individuals, the population risk is calculated by
either summing the risks for each individual or by multiplying the average risk
across individuals by the population size. The total population risk may be
expressed as
total population risk = average individual risk for group a x number (2-7)
of people in group a + average individual risk for group b x number
of people in group b + average individual risk for group n x number
of people in group n.
Likewise, when multiple contaminant exposures occur, the total risk will equal the
sum of the risks from individual contaminants at each exposure level.
2.5.2 Noncarcinogenic Toxicity
For chronic systemic toxicants, the RfD is used as a reference point in assessing
risk. The RfD is an estimate, with an uncertainty of perhaps an order of
magnitude, of a daily exposure that is likely to be without appreciable risk of
deleterious health effects in the human population (including sensitive subgroups)
over a lifetime.
2.5.2.1 Individual Risk—
The comparison of exposure to the RfD indicates the degree to which exposure is
greater or less than the RfD. The following equation expresses this relationship:
ratio = exposure/RfD (2-8)
where
exposure = total exposure to a single contaminant from all sources
(mg/kg-d)
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2. RISK ASSESSMENT METHODS
RfD = reference dose or other noncarcinogenic exposure limit.
When the ratio obtained in the above equation is equal to or greater than 1 (i.e.,
when exposure exceeds the RfD), the exposed populations may be at risk.
Although a margin of safety is incorporated into RfDs (see Section 2.3), actual
thresholds are usually not known. Consequently, exposure above the RfD is not
recommended. The likelihood of risk is related to the degree to which exposure
exceeds the RfD. Risk also depends on individual characteristics; susceptibility to
toxic exposures varies considerably in most populations. Consequently, the
primary use of RfDs is to provide a protective exposure limit rather than to predict
risks. In practice, however, they are often used to estimate risk.
2.5.2.2 Population Risk—
The population risk is expressed as the number of individuals with exposure levels
greater than the RfD:
noncarcinogenic risk = population with exposure greater than the RfD. (2-9)
Reviewing the health basis for the risk estimate is useful when evaluating the risk
estimates. A wide range of effects is used to establish RfDs. Some are very
serious (e.g., retarded growth, liver damage, infertility, brain dysfunction) and
others are of less concern (e.g., changes in enzyme levels indicative of preliminary
stages of toxicity). In most cases the less serious effects will lead to serious effects
as exposure levels increase above the RfD. This type of toxicity information should
be considered when reviewing risk estimates.
Nonfish sources of exposure may be an important contributor to overall
exposure. In some cases, exposure to a contaminant via fish consumption
alone may not generate risk at the population's consumption level, but
exposure to the contaminant in fish and other foods, water, soil, or air may
exceed the RfD. Total exposure information can be used to obtain a much more
accurate assessment of risk. When exposure occurs via other sources, the lack of
total exposure assessment leads to an underestimate of exposure, and potentially
of risk. Accurate risk information provides a more appropriate basis for decisions
concerning the need for fish advisories.
An alternative approach is to express the dose as the magnitude by which the
NOAEL exceeds the estimated dose (termed the margin of exposure, or the MOE).
Where the MOE is greater than the product of the uncertainty and modifying
factors (used in calculating an RfD from a NOAEL), then concern is considered to
be low (Barnes and Dourson, 1988).
2.5.3 Subpopulation Considerations
A major goal in evaluating population risks is the identification of target populations.
This document defines target populations as fish consumers determined by
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2. RISK ASSESSMENT METHODS
decisionmakers to be in need offish advisory programs. This section discusses the
criteria for such a decision.
The identification of target populations involves both risk assessors and risk
managers and requires both scientific and policy judgments.
A population would usually be targeted because they consume fish containing
contaminants that may pose health hazards. In some cases, they may have known
high exposures; in other cases, state agencies may have limited information
suggesting they are at risk. Regardless of the supporting data available,
determining who the target populations are is a critical step in establishing a fish
advisory program.
A risk-based approach can be used to identify target populations. This approach
requires decisions concerning the level of "acceptable" risk for carcinogenic and
noncarcinogenic effects. For example, a health agency may determine that any
population with cancer risk levels greater than 1 in 1 million requires a con-
sumption advisory. For noncarcinogenic effects, exposures greater than the RfD
by a factor of 1,10, or some other value may be chosen to determine which groups
require protection under a fish advisory program. Establishing an exposure limit for
the purposes of identifying at-risk populations enables state agencies to equitably
screen populations to determine where action is needed. Different subgroups
within a population will often have differing consumption rates and may need to be
considered individually to adequately address their levels of risk and need for
program assistance. For example, children consuming contaminated fish at a rate
that is safe for adults may be at risk due to their small body size and increased
intake per unit of body weight (mg/kg-d). Choosing the levels at which populations
are determined to need such advisories is a policy decision.
Defining acceptable risk has been a difficult problem at both the federal and local
level. Federal programs have targeted various levels of cancer risk in developing
regulations and guidance, and these levels often change over time and may be
modified based on the needs of particular areas. "Acceptable" risk has also been
defined and redefined in a number of legal cases.
Decisions concerning acceptable risk levels are often considered high-level policy
decisions because they may affect the public's health directly. Many states have
specific guidance written into their legislation concerning benchmark levels of risk
(e.g., 1 in 1 million cancer risk is targeted in New Jersey for drinking water
contaminants, modified by feasibility considerations).
Because of the importance of decisions concerning acceptable risk levels,
state agencies are encouraged to seek input from a variety of sources,
including target populations, when establishing these levels. The selection
of specific groups as target populations is a critical decision because it
affects who will be served, the levels of potential risk of those who will not
be served, and the scope of the fish advisory program needed. EPA
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encourages state, local, and tribal governments to consider the most
sensitive populations when establishing programs. "Sensitive" in this
context means those people who are at greatest risk due to their exposure,
age, predisposing conditions, or other factors.
Some population groups may warrant more restrictive risk levels (e.g., children
may be considered more susceptible than some other subgroups); however, levels
of protection and provisions of services should be equitable across all persons
served.
2.5.4 Multiple Species and Multiple Contaminant Considerations
Readers are encouraged to take multiple species consumption and/or multiple
contaminant exposures into account when developing consumption limits and/or
assessing risk. Methods for doing so are described in Sections 2.4.5.4, 3.4, and
3.5.
2.5.5 Incorporating Considerations of Uncertainty in Consumption Limits
Previous sections have discussed the many uncertainties associated with the
estimates of exposure and toxicity data assessments that form the basis of the risk
assessment and the derivation of risk-based consumption limits. Readers may
wish to estimate the direction the uncertainties are likely to have on the risk
estimates (i.e., do these uncertainties tend to exaggerate or diminish potential risk).
The assumptions made in the risk assessments to account for uncertainties need
to be clearly outlined (e.g., Section 2.3.5 contains a description of the nature of the
uncertainties associated with each uncertainty factor applied in deriving an RfD).
The use of the 95 percent upper confidence limit for the slope of the dose-
response function at low doses for carcinogens is an example of a conservative
assumption imbedded in most cancer slope factors. Likewise, exposure assess-
ments frequently include conservative assumptions where data on actual exposure
are absent, such as the assumption that no dose modification occurs when the
cooking and preparation methods of target populations are unknown. Where
possible, readers are encouraged to attempt to quantify the magnitude of the effect
of such assumptions on the numerical risk estimates.
2.6 SUMMARIZING RISK DATA
This section describes methods for summarizing population exposure and risk. The
risk assessment process can generate considerable data on various populations
and geographic areas with details on numerous contaminants and levels of
exposure. Organization of these data is useful so that the results can be reviewed
in a meaningful way. Because different circumstances will require different data
arrays, a number of templates are provided (Tables 2-5, 2-6, and 2-7) for
organizing risk information for various purposes.
2-55
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2. RISK ASSESSMENT METHODS
The presentation of the templates proceeds from the most specific (risk levels for
a specific population at a specific waterbody) to more general risk summaries for
a large geographic area. The templates are offered as a convenience and may
contain entry areas that are not appropriate for all circumstances. State agency
staff are encouraged to modify these or omit areas as needed.
Table 2-5 is a template that can be used to organize exposure data, risk values,
and risk estimates. It is designed to be used for a specific population in a specific
location with exposure to a contaminant at a known level. This table provides entry
areas for the various factors that are used in calculating risk, as well as the actual
risk estimates. Depending on the type of contaminants present and population
characteristics, estimating risks for various subgroups may be advisable. This data
display will allow agencies to highlight which groups within a population are at
highest risk and to summarize the risks to a particular population. This table can
also be used to evaluate the varied impacts on risk that may occur as a result of
changing assumptions concerning consumption patterns, contaminant concentra-
tions, and risk values.
Fish contaminants and contaminant concentrations are listed in the left column. If
different concentrations are expected in different size fish, different tables can be
developed for the various concentrations. Table 2-5 includes entries for central
tendencies, high-end, and bounding exposure and risk estimates. It is not expected
that all these variables will be calculated for all groups and conditions. This
information, however, provides a range of estimates that can be used in prioritizing
activities and designing appropriate programs. The template has entry areas for
both fish and nonfish exposures.
Some agencies may not have information on nonfish exposures or may choose not
to evaluate other sources of exposure in determining appropriate fish advisories.
Risk assessors may modify the categories of information listed in this table to suit
the specific characteristics of their local populations and fish advisory programs.
Table 2-5 also provides information lines for risks to women 18 to 45 years of age,
the reproductive age for many women. This separate entry area was provided
because many health officials are particularly concerned about developmental
effects that may arise from exposure to long-term or bolus doses of fish
contaminants, especially mercury. Separate entry areas for children were also
provided because their consumption in relation to their body weight is often greater
than that of adults. Consequently, their risks may be higher for noncarcinogens
(carcinogenic risk estimates are based on a lifetime exposure, including childhood).
Evaluation of the risks to multiple groups may be warranted when more than one
population uses a particular waterbody. Under those circumstances, various data
summaries may be needed to provide data for differing fish advisories. For
example, sport fishers and subsistence fishers may use the same waterbody but
have different risks based on their varied consumption habits.
2-56
-------
Table 2-5. Risk Estimates
Location:
Population:
Population Size:
Contaminant:
Contaminant Concentration:
Specific
Subgroups
Total Population
<18yr
>18yr
Women, 18-45
Fish Exposure
Estimates
Central
Tendency
High-End
Estimate9
Other
Exposures
Subtotal of Other
Exposures
Central
Tendency
High-End
Estimate
Total All Exposures
Central
Tendency
High-End
Estimate
Risk Values
Non-
carcinogen
Carcinoge
n
Alternatives
Other Factors (e.g.,
special
susceptibilities due
to nutritional
status, disease,
etc.)
10
33
CO
7;
m
m
m
O
o
Risk Estimate
Central Tendency
Noncarcinogen
(% of Rf D)
Fish
Only
All
Exposures
Carcinogen
(Lifetime Risk)
Fish
Only
All
Exposures
Alternatives
(% of Alternatives)
Fish
Only
All
Exposures
High-End Estimate
Noncarcinogen
(% of Rf D)
Fish
Only
All
Exposures
Carcinogen
(Lifetime Risk)
Fish
Only
All
Exposures
Alternatives
(% of Alternatives)
Fish
Only
All
Exposures
-------
Table 2-6. Risk Characterization
Location:
Population:
Population Size:
Contaminant
Level (mg/kg)
Total
Central Tendency
Carcinogen
(Lifetime Risk)
Noncarcinogen
(% of RfD)
Alternatives
(%ofAltern.)
High-End Estimate or Bounding Estimate
Carcinogen
(Lifetime Risk)
Noncarcinogen
(% of RfD)
Alternatives
(%ofAltern.)
en
oo
10
33
CO
7;
m
m
m
O
o
-------
2. RISK ASSESSMENT METHODS
Table 2-7. Risk Summaries for a Waterbody
Risk Estimates Based on High-End Exposures
Population Group Cancer Risks Noncancer Risks Other Risks
Total Population A
<18yr
>18yr
Women 18-45 yr
Total Population B
<18yr
>18yr
Women 18-45 yr
Total Population C
<18yr
>18yr
Women 18-45 yr
Aggregate of A,B,C
<18yr
>18yr
Women 18-45 yr
Table 2-5 provides entry areas for the various factors used to calculate risk. State
agencies may wish to use this format to evaluate the sensitivity of the final risk
estimates to variations in input factors such as fish exposure, other exposures, risk
values, contaminant concentrations, and body weight. This type of sensitivity
analysis will provide information on the importance of the various factors. When
uncertainty exists about one of the inputs, such as a risk value or contaminant
level, its relative importance in the overall estimates of risk can be evaluated.
Table 2-6 provides a template to be used to summarize risk data for a specific
population using information presented in Table 2-5. This table focuses on health
risk assessment and does not include information on the variables used to
calculate risk, such as exposures and risk values. Table 2-6 is particularly useful
when the same populations are exposed to more than one contaminant or multiple
2-59
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2. RISK ASSESSMENT METHODS
concentrations of the same contaminant. The risk results for different contaminants
may be entered by listing different chemicals down the left column and their
corresponding risks across the same row. Alternatively, risks resulting from
different contaminant levels can be entered in the left column when exposures to
varied species are occurring with differing concentrations of contaminants.
If an additive effect is suspected, the total carcinogenic or noncarcinogenic risks
could then be summed for the population or subgroup. Risk estimates may be
modified if either a synergistic or antagonistic effect is expected.
Table 2-7 is a template designed to summarize risks for more than one population
using a particular waterbody. This approach allows state agencies to obtain an
overall estimate of the risks associated with fishing in a specific waterbody. This
type of information may be particularly useful in evaluating the need for an advisory
over a large geographic area and for a number of waterbodies.
Geographically based fish advisory efforts may target particular regions or areas
based on overall risks for the waterbodies in an area. Waterbody-specific risk data
can be used to prioritize efforts and may show concentrations of risk that would not
be obvious using small population units as groups for comparison. They may also
be used to determine that no action is necessary if the sum of all population risks
is negligible. If a geographic approach is used in the development of fish
advisories, Section 6, which gives an overview of mapping techniques, should be
consulted.
Table 2-7 uses summary information from Tables 2-5 or 2-6 and assumes that
state agencies will have focused their attention on a particular aspect of the risk
distribution (i.e., central tendency, high-end, or bounding estimates). High-end
values are listed in the table because it is recommended that fish advisories be
based on highly, but realistically, exposed individuals and risks. State agencies
may elect, however, to choose some other portion of the risk distribution.
Table 2-7 also provides data entry areas for three populations surrounding a water-
body (A, B, and C) and for various subgroups within those areas. Data entry areas
are provided for cancer, noncancer, and "other" risks. The third variable is provided
because some decisionmakers may wish to evaluate more than one type of risk
in a particular category or use more than one risk value (e.g., liver damage and
developmental toxicity). Data entry areas are also provided at the bottom of the
table to summarize the risks across populations for the total population and for
various subgroups. As with all the tables in this document, state agencies may
wish to modify this table to address their specific needs.
State agencies may wish to compare risks at different waterbodies over large
geographic areas. Table 2-8 provides a template designed to summarize risk data
collected for specific waterbodies and populations. The table may be used to
summarize risks to the overall populations or to specific subpopulations using a
2-60
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2. RISK ASSESSMENT METHODS
waterbody. If subpopulation risks are of interest, the format provided in Table 2-8
can be followed with four rows used for each waterbody.
Table 2-8. Risk Summaries for a Geographic Area
Waterbody Location
Total Risk:
Risk Estimates Based on
High-End Exposures
Carcinogenic
Effects
Noncarcinogenic
Effects
2-61
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3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
SECTION 3
DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
3.1 OVERVIEW AND SECTION ORGANIZATION
This section describes the derivation and use of the risk-based consumption limit
tables provided in Section 4. Consumption limit tables were developed for each
of the 25 target analytes listed in Table 1-1 and described in further detail in
Volume 1 of this series. This section discusses
Equations used to calculate the consumption limit tables
Default values used in developing the consumption limit tables
Modifications to the consumption limit calculations to allow for different input
values and for multiple species consumption and/or multiple contaminant
exposure.
Methods for deriving consumption limits for chemical contaminants with carcino-
genic and/or noncarcinogenic effects are described. When available data
indicate that a target analyte is associated with both carcinogenic and
noncarcinogenic health effects, consumption limits based on both types of effects
are calculated. In these cases, it is recommended that the toxicological effect
resulting in the more conservative consumption limits be used to issue an
advisory since resulting limits would be protective of both types of health effects.
Methods for calculating consumption limits for a single contaminant in a multiple
species diet or for multiple contaminants causing the same chronic health effects
endpoints are also discussed. Species-specific consumption limits are calculated
as fish meals per month, at various fish tissue concentrations, for noncancer and
cancer health endpoints.
Developing fish consumption limits also requires making assumptions about the
edible portions of fish because most chemical contaminants are not evenly
distributed throughout the fish. The portion of the fish typically eaten may vary by
fish species and/or the dietary habits of the fisher population of concern. Most
fishers in the United States consume fish fillets. Therefore, it is recommended
that contaminant concentrations be measured using skin-on fillets for scaled fish
species and skinless fillets for scaleless fish species (e.g., catfish) (see Section
6.1.1.6 in Volume 1 of this series for further discussion of edible fish and shellfish
sample types). However, for populations that ingest whole fish, consumption
3-1
-------
3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
values corresponding to whole fish contaminant concentrations are more appro-
priate. Fish consumption patterns are discussed in more detail in Appendix B.
People may be exposed to one or more fish contaminants through sources or
pathways other than through consumption of recreationally or subsistence caught
fish. These sources include ingestion of contaminated commercially caught fish,
other contaminated foods, or contaminated drinking water; inhalation of the con-
taminant; or dermal contact with contaminated materials including soil and sedi-
ment. Caution should be used in setting health safety standards that do not take
these other sources into account (see Section 2 for further discussion). Methods
for quantifying exposure via sources other than consumption of recreationally or
subsistence caught fish are not discussed in detail in this series.
3.2 EQUATIONS USED TO DEVELOP RISK-BASED CONSUMPTION LIMITS
Two equations are required to derive meal consumption limits for either carcino-
genic or noncarcinogenic health effects. The first equation (3-1 for carcinogenic
effects or Equation 3-3 for noncarcinogenic effects) is used to calculate daily
consumption limits in units of milligrams of edible fish per kilogram of consumer
body weight per day (mg/kg-d); the second equation (3-2) is used to convert daily
consumption limits to meal consumption limits over a specified period of time
(e.g., 1 month). Toxicological benchmark values for carcinogenic and non-
carcinogenic health effects used in the calculation of risk-based consumption
limits are summarized in Table 3-1.
3.2.1 Calculation of Consumption Limits for Carcinogenic Effects
To calculate consumption limits for carcinogenic effects, it is necessary to specify
an "acceptable" lifetime risk level (ARL). The appropriate risk level for a given
population is determined by risk managers; see Volume 3 for further discussion
of selection of appropriate risk level. This document presents consumption limits
that were calculated using a risk level of 1 in 100,000 (10~5). Equations 3-1 and
3-2 were used to calculate risk-based consumption limits for the 12 target
analytes with cancer slope factors (see Table 3-1), based on an assumed 70-yr
exposure. A 70-yr lifetime is used in keeping with the default value provided in
EPA's Exposure Factors Handbook (U.S. EPA, 1990a). This is a normative value;
individuals may actually be exposed for greater or lesser periods of time,
depending on their lifespan, consumption habits, and residence location. It
should be noted that no populations were identified as being particularly
susceptible to the carcinogenic effects of the target analytes.
3.2.1.1 Calculation of Daily Consumption Limits-
Equation 3-1 calculates an allowable daily consumption of contaminated fish
based on a contaminant's carcinogenicity, expressed in kilograms of fish
consumed per day:
3-2
-------
3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
Table 3-1. Risk Values Used in Risk-Based Consumption Limit Tables
Noncarcinogens Carcinogens
Target Analyte
Metals
Arsenic (inorganic)0
Cadmium
Mercury (methylmercury)d
Selenium
Tributyltinb
Organochlorine Pesticides
Total chlordane (sum of c/s- and trans-
chlordane,
c/s- and frans-nonachlor, and oxychlordane)6
Total DDT (sum of 4,4'- and 2,4'-
isomers of DDT, DDE, and DDD)f
Dicofol9
Dieldrin
Endosulfan (I and II)
Endrin
Heptachlor epoxide
Hexachlorobenzene
Lindane (y-hexachlorocyclohexane; y-HCH)'
Mirex
ToxaphenehJ
Organophosphate Pesticides
Chlorpyrifosk
Diazinon'
Disulfoton
Ethion
Terbufosm
Chlorophenoxy Herbicides
Oxyfluorfen"
PAHs0
PCBs
Total PCBs
Dioxins/furansq
CSF = Cancer slope factor (mg/kg-d)"1.
DDD = p,p' -dichlorodiphenyldichloroethane.
DDE = p,p' -dichlorodiphenyldichloroethylene
DDT = p,p' -dichlorodiphenyltrichloroethane.
NA = Not available in EPA's Integrated Risk
Information System (IRIS, 1999).
Chronic RfDa
(mg/kg-d)
3x irj-4
1 x irj-3
1 x irj-4
5x 1Q-3
3x 1Q-4
5x 1Q-4
5x irj-4
4x 1Q-4
5x 10-5
6x 1Q-3
3x 1Q-4
1.3x 10 5
8x 1Q-4
3x 1Q-4
2x 1Q-4
2.5 x 1Q-4
3x 1Q-4
7x 1Q-4
4x 1Q-5
5x 1Q-4
2x 1Q-5
3x 1Q-3
NA
2x 1Q-5
NA
PAH =
PCB =
RfD
CSFa
(mg/kg-d)1
1.5
NA
NA
NA
NA
0.35
0.34
withdrawn
16
NA
NA
9.1
1.6
1.3
NA
1.1
NA
NA
NA
NA
NA
7.32 x 1Q-2
7.3
2.0p
1.56x 105
Polycyclic aromatic hydrocarbon.
Polychlorinated biphenyl.
= Oral reference dose (mg/kg-d).
(continued)
3-3
-------
3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
Table 3-1 (continued)
3 Unless otherwise noted, values listed are the most current oral RfDs and CSFs in EPA's IRIS database
(IRIS, 1999).
b The RfD value listed is for the IRIS (1999) value for tributyltin oxide.
c Total inorganic arsenic should be determined.
d Because most mercury in fish and shellfish tissue is present primarily as methylmercury (NAS, 1991;
Tollefson, 1989) and because of the relatively high cost of analyzing for methylmercury, it is recommended
that total mercury be analyzed and the conservative assumption be made that all mercury is present as
methylmercury. This approach is deemed to be most protective of human health and most cost-effective.
The National Academy of Sciences (NAS) conducted an independent assessment of the RfD and
concluded, "On the basis of its evaluation, the committee consensus is that the value of EPA's current RfD
for methylmercury, 0.1 ug/kg per day, is a scientifically justifiable level for the protection of human health."
e The RfD and CSF values listed are derived from studies using technical-grade chlordane (IRIS, 1999). No
RfD or CSF values are given in IRIS (1999) for the c/s- and frans-chlordane isomers or the major chlordane
metabolite, oxychlordane, or for the chlordane impurities c/s- and frans-nonachlor. It is recommended that
the total concentration of c/s- and frans-chlordane, c/s- and frans-nonachlor, and oxychlordane be
determined.
f The RfD value listed is for DDT. The CSF value is 0.34 for total DDT (sum of DDT, DDE, and ODD). The
CSF value for DDD is 0.24. It is recommended that the total concentration of the 2,4'- and 4,4'-isomers
of DDT and its metabolites, DDE and DDD, be determined.
9 The RfD value is from the Registration Eligibility Decision (RED). Dicofol (U.S. EPA, 1998a).
h The RfD value listed is from the Office of Pesticide Program's Reference Dose Tracking Report (U.S. EPA,
1997c).
1 IRIS (1999) has not provided a CSF for lindane. The CSF value listed for lindane is from HEAST, 1997.
1 The RfD value has been agreed upon by the Office of Pesticide Programs and the Office of Water.
K Because of the potential for adverse neurological developmental effects, EPA recommends the use of a
Population Adjusted Dose (PAD) of 3x10"5 mg/k-d for infants, children to the age of six, and women ages
13-50 (U.S. EPA, 2000b).
' The RfD value is from a memo data April 1, 1998, Diazinon: Report of the Hazard Identification
Assessment Review Committee. HED DOC. NO. 012558 (U.S. EPA, 1998c).
m The RfD value listed is from a memorandum dated September 25, 1997; Terbufos-FQPA Requirement
Report of the Hazard Identification Review (U.S. EPA, 1997h).
" The CSF value is from a memo dated 9/24/98; REVISED Oxyfluorfen (GOAL) Quantitative Risk
Assessment (Q1*) Based on CD-1 Male Mouse Dietary Study With 3/4's Interspecies Scaling Factor. HED
Document No. 012879 (U.S. EPA, 1998c).
0 The CSF value listed is for benzo[a]pyrene. Values for other PAHs are not currently available in IRIS
(1999). It is recommended that tissue samples be analyzed for benzo[a]pyrene and 14 other PAHs and
that the order-of-magnitude relative potencies given for these PAHs (Nisbet and LaGoy, 1992; U.S. EPA,
1993b) be used to calculate a potency equivalency concentration (PEC) for each sample (see Section
5.3.2.4 of Volume 1).
p The CSF is based on a carcinogenicity assessment of Aroclors 1260, 1254, 1242, and 1016. The CSF
presented is the upper-bound slope factor for food chain exposure. The central estimate is 1.0 (IRIS,
1999).
q The CSF value listed is for 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) (HEAST, 1997). It is recommended
that the 17 2,3,7,8-substituted tetra- through octa-chlorinated dibenzo-p-dioxins and dibenzofurans and
the 12 dioxin-like PCBs be determined and a toxicity-weighted total concentration be calculated for each
sample, using the method for estimating Toxicity Equivalency Concentrations (TEQs) (Van den Berg et
al., 1998).
3-4
-------
3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
_ ARL • BW
llm = CSF.C (3-D
m
where
CR|im = maximum allowable fish consumption rate (kg/d)
ARL = maximum acceptable individual lifetime risk level (unitless)
BW = consumer body weight (kg)
CSF = cancer slope factor, usually the upper 95 percent confidence limit
on the linear term in the multistage model used by EPA [(mg/
kg-d)~1], (see Section 2 for a discussion of this value)
Cm = measured concentration of chemical contaminant m in a given
species offish (mg/kg).
The calculated daily consumption limit (CRHm) represents the amount of fish (in
kilograms) expected to generate a risk no greater than the maximum ARL used,
based on a lifetime of daily consumption at that consumption limit.
3.2.1.2 Calculation of Meal Consumption Limits —
Daily consumption limits may be more conveniently expressed as the allowable
number of fish meals of a specified meal size that may be consumed over a
given time period. The consumption limit is determined in part by the size of the
meal consumed. An 8-oz (0.227-kg) meal size was assumed. Equations 3-1 and
3-2 can be used to convert daily consumption limits, the number of allowable
kilograms per day (calculated using Equation 3-1), to the number of allowable
meals per month:
• 1 /o r)\
m ap \3-t-)
MS
where
Crmm = maximum allowable fish consumption rate (meals/mo)
Cr,im = maximum allowable fish consumption rate (kg/d)
MS = meal size (0.227 kg fish/meal)
Tap = time averaging period (365.25 d/12 mo = 30.44 d/mo).
Equation 3-2 was used to convert daily consumption limits, in kilograms, to meal
consumption limits over a given time period (month) as a function of meal size.
Monthly consumption limits for carcinogenic effects in adults in the general
population were derived for 13 of the 25 target analytes in Section 4.
Other consumption rates, such as meals per week, could also be calculated
using this equation by substituting, for example, 7 d/wk for 30.44 d/mo. In using
3-5
-------
3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
Equation 3-2 in the table calculations in Section 4, the reader should note that 1
month was expressed as 365.25 d/12 mo or 30.44 d/mo.
3.2.1.3 Input Parameters-
Calculating risk-based consumption limits for carcinogenic effects requires
developing appropriate values for the parameters in the equations. The default
values used to calculate the consumption limits listed in Section 4 are shown in
Table 3-2; a range of values is provided for the measured contaminant
concentration in fish tissue (CJ to represent a broad spectrum of contaminant
concentrations. See consumption limit tables in Section 4. Development and
modification of these values are discussed in Section 3.3.
EXAMPLE 1: Calculating Monthly Consumption Limits for
Carcinogenic Health Endpoints in the General Population
for Chlordane
Table 3-2. Input Parameters for Use in Risk Equations
Equation Parameter3
Values
Maximum acceptable risk level (ARL)
Cancer slope factor (CSF)b
Reference dose (RfD)
Consumer body weight (BW)
Average fish meal size (MS)
Measured contaminant concentration
in edible fish and shellfish tissue (Cm)c
Time-averaging period (T )
10-5(unitless)
(mg/kg-d)-1
mg/kg-d
70 kg (general adult population)
8 oz (0.227 kg)
mg/kg (ppm)
varies with local conditions for each
chemical contaminant, for each
species, and for each size (age) class
within a species
30.44 d/mo (monthly limit)
Selection of the appropriate maximum acceptable risk level, consumer body weight, and
average fish meal size are considered risk management decisions. For information
regarding these values, see Sections 2 and 5 of this document and Volume 3.
Most of the CSFs and RfDs were obtained from EPA's Integrated Risk Information
System (IRIS, 1999). The RfDs not listed in IRIS were obtained from EPA's Office of
Pesticide Programs. The CSFs and RfDs used in the risk equations are listed in Table 3-1
and are discussed in Section 5.
Values for contaminant concentrations should be determined from local fish sampling
and analysis programs conducted in the waterbody of concern as described in Volume
1.
3-6
-------
3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
Using Equations 3-1 and 3-2, the monthly meal consumption limits were cal-
culated for the carcinogenic effects of chlordane for adults in the general
population as shown in Table 3-3. Note: In this section, the monthly
consumption limits for chlordane for both carcinogenic and chronic
(noncarcinogenic) health effects are used to illustrate various modifications
to the monthly consumption limit tables.
3.2.2 Calculation of Consumption Limits for Noncarcinogenic Effects
Noncarcinogenic health effects caused by consumption of contaminated fish
include systemic effects such as liver, kidney, neurological, muscular, ocular,
reproductive, respiratory, circulatory, or other organ toxicities and adverse
developmental/reproductive effects from acute and chronic exposure. Risk-based
consumption limit tables for chronic exposure health effects were developed for
adults and young children for 23 of the 25 target analytes using RfDs for chronic
systemic health effects.
3.2.2.1 Calculation of Daily Consumption Limits—
Equation 3-3 calculates an allowable daily consumption (CRHm) of contaminated
fish, based on a contaminant's noncarcinogenic health effects, and is expressed
in kilograms of fish per day:
„„ RfD • BW
CRiim = F (3-3)
where
CR|im = maximum allowable fish consumption rate (kg/d)
RfD = reference dose (mg/kg-d)
BW = consumer body weight (kg)
Cm = measured concentration of chemical contaminant m in a given
species offish (mg/kg).
CR|im represents the maximum lifetime daily consumption rate (in kilograms of
fish) that would not be expected to cause adverse noncarcinogenic health
effects. Most RfDs are based on chronic exposure studies (or subchronic studies
used with an additional uncertainty factor). Because the contaminant
concentrations required to produce chronic health effects are generally lower
than those causing acute health effects, the use of chronic RfDs in developing
consumption limits is expected to also protect consumers against acute health
effects. They are designed to protect the most sensitive individuals.
To calculate weekly fish meal consumption limits, Equation 3-3 was modified as
follows:
3-7
-------
3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
_, RfD x BW
= ~~ (3'4)
Using this equation, one can calculate the level of chemical contamination (CJ
in a given species offish assuming that a 70-kg adult consumes a maximum of
one 8-oz (0.227-kg) meal/wk.
3.2.2.2 Calculation of Meal Consumption Limits—
Equation 3-2 is used to convert daily consumption limits, in kilograms, to meal
consumption limits over given time periods as a function of meal size. An 8-oz
meal size was assumed in the calculations. Monthly consumption limits were
derived for all target analytes in Section 4 except PAHs and dioxins, for which
RfD values are not available. Monthly consumption limits pertain to recreational
fishers (see Section 2.4.5.4). Where appropriate, risk assessors may choose to
derive consumption limits based on a shortertime-averaging period such as a 14-
d period (see Section 3.3.6). Note that, irrespective of the time-averaging period
selected (e.g., 7-d, 10-d, 14-d, monthly), the same chronic systemic RfDs are
applicable; the difference is in the averaging periods used in Equation 3-2.
Note: This approach does not expressly limit the amount of fish that may be
consumed in a given day during the specified time period, so care must be taken
to inform consumers of the dangers of eating large amounts of contaminated fish
in one meal when certain acute or developmental toxicants are of concern.
3.2.2.3 Input Parameters—
For noncarcinogenic effects, calculating risk-based consumption limits requires
developing appropriate values for similar parameters to those required for
carcinogenic effects (see Table 3-2).
3.2.3 Developmental Effects
This guidance document does not calculate consumption limits specifically for
developmental effects. For the majority of target analytes, sufficiently detailed
developmental toxicity data are not available. For two analytes, methylmercury
and PCBs, sufficient data are available demonstrating that women exposed to
these chemicals may transfer sufficient amounts in utero or through breast
feeding to induce pre- or postnatal developmental damage in their offspring. The
interim RfD for methylmercury (1 x 10~4 mg/kg-d) is based on developmental
effects in humans (i.e., neurologic changes in Iraqi children who had been
exposed in utero).
3-8
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3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
EXAMPLE 2: Calculating Monthly Consumption Limits for Chronic
Systemic Health Endpoints for Recreational Fishers for
Chlordane
Using Equations 3-3 and 3-2, the monthly meal consumption limits were
calculated for the noncarcinogenic and carcinogenic health effects of
chlordane for recreational fishers as shown in Table 3-3. Note: In comparing
the consumption limit tables for chlordane based on carcinogenic and
noncarcinogenic effects for the general population, it is apparent that the
carcinogenic endpoint results in a more conservative consumption limit
assuming an ARL of 10~5 and equivalent meal sizes and contaminant
concentrations in fish tissues. For example, based on a chemical contaminant
level in fish tissue of 0.1 ppm, an adult could eat seven 8-oz fish meals
assuming an ARL of 10~5. Given the same level of tissue contamination, an
adult could eat >30 8-oz meals per month based on noncarcinogenic effects
of chlordane. To protect consumers from both the carcinogenic and
noncarcinogenic effects of chlordane, a risk assessor may choose to base
consumption limits on the more conservative meal sizes derived for
carcinogenic effects. In this situation, a risk assessor or risk manager may wish
to issue the consumption advisory based on the carcinogenic effects of
chlordane, which would be protective of chronic health effects given the
above-stated assumptions.
Thus, the consumption limits would be protective against developmental effects
for methylmercury.
3.3 DEFAULT AND ALTERNATIVE VALUES FOR CALCULATING CONSUMPTION LIMITS
The consumption limit tables provided in Section 4 are based on default values
for consumer body weights and average meal sizes. This section describes the
default values shown in Tables 3-1 and 3-2 and provides alternative input values
and multipliers for use in modifying and/or recalculating the consumption limit
tables.
Seven variables are involved in calculating the values in the consumption limit
tables (see Equations 3-1 through 3-3):
Maximum acceptable risk level (ARL)
Cancer slope factor (CSF)
Chronic reference dose (RfD)
Consumer body weight (BW)
Fish meal size (MS)
Contaminant concentration in edible fish tissue (CJ
Time-averaging period (30-d period).
3-9
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3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
Monthly meal consumption limit tables for both the carcinogenic and noncarcin-
ogenic health effects of chlordane are used as examples to illustrate the effects
of modifying one or more of the variables listed above.
Table 3-3. Monthly Fish Consumption Limits for Carcinogenic and Noncarcinogenic
Health Endpoints - Chlordane
Risk Based Consumption Limit3
Fish Meals/Month
Unrestricted (>16)
16
12
8
4
3
2
1
0.5
None (<0.5)
Noncancer Health Endpointsb
Fish Tissue Concentrations
(ppm, wet weight)
0-0.15
>0.15-0.29
>0.29-0.39
>0.39-0.59
>0.59-1.2
>1.2- 1.6
>1.6-2.3
>2.3-4.7
>4.7 - 9.4
>9.4
Cancer Health Endpoints0
Fish Tissue Concentrations
(ppm, wet weight)
0 - 0.0084
>0.0084- 0.017
>0.017-0.022
>0.022 - 0.034
>0.034 - 0.067
>0.067 - 0.089
>0.089-0.13
>0.13-0.27
>0.27 - 0.54
>0.54
a The assumed meal size is 8 oz (0.227 kg). The ranges of chemical concentrations presented are conservative, e.g., the
12-meal-per-month levels represent the concentrations associated with 12 to 15.9 meals.
b Chronic, systemic effects.
c Cancer values represent tissue concentrations at a 1 in 100,000 risk level.
Notes:
1. Consumption limits are based on an adult body weight of 70 kg, an RfD of 5x10"4 mg/kg-d, and a cancer slope factor
(CSF)of0.35(mg/kg-d)-1
2. None = No consumption recommended.
3. In cases where >16 meals per month are consumed, refer to Equations 3-1 and 3-2, Section 3.2.1.2, for methods to
determine safe consumption limits.
4. The detection limit for chlordane is 5 x 10"3 mg/kg.
5. Instructions for modifying the variables in this table are found in Section 3.3.
6. Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD). When the monthly limit
is consumed in less than 1 month (e.g., in a few large meals), the daily dose may exceed the RfD (see Section 2.3).
3.3.1 Maximum Acceptable Risk Level
The consumption limit tables shown in Section 4 for target analytes with carcino-
genic effects were calculated for maximum individual ARL of 10~5. Note that the
variable ARL appears in the numerator of Equation 3-1, the equation for
calculating the daily consumption limit for carcinogens. Because ARL appears in
multiples of 10, one may derive new meal consumption limits from the existing
tables by multiplying or dividing the existing meal consumption limits by factors
3-10
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3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
of 10, as appropriate. In the same way, changing the ARL by a factor of 10 would
cause the same meal consumption limits to be valid for chemical concentrations
10 times higher or 10 times lower than those associated with the original ARL
(see Table 3-4).
3-11
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3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
Table 3-4. Monthly Fish Consumption Limits for Carcinogenic Health Endpoints - Chlordane
Risk Based
Consumption
Limit3
Recommended Risk-Based Consumption Limit
(meals per month, 8-oz meal size)
Fish tissue Concentrations (ppm, wet weight)
Fish Meals/Month
Unrestricted (>16)
16
12
8
4
3
2
1
0.5
None (<0.5)
ARLKT1
0 - 0.084
>0.084-0.17
>0.17-0.22
>0.22 - 0.34
>0.34 - 0.67
>0.67-0.89
>0.89-1.3
>1.3-2.7
>2.7 - 5.4
>5.4
ARL105
0 - 0.0084
>0.0084- 0.017
>0.017-0.022
>0.022 - 0.034
>0.034 - 0.067
>0.067 - 0.089
>0.089-0.13
>0.13-0.27
>0.27 - 0.54
>0.54
ARL106
0 - 0.00084
>0.00084-0.0017
>0.001 7 -0.0022
>0.0022 - 0.0034
>0.0034 - 0.0067
>0.0067 - 0.0089
>0.0089- 0.013
>0.013-0.027
>0.027 - 0.054
>0.054
ARL107
0 - 0.000084
>0.000084- 0.0001 7
>0.0001 7 -0.00022
>0.00022 - 0.00034
>0.00034 - 0.00067
>0.00067 - 0.00089
>0.00089-0.0013
>0.001 3 -0.0027
>0.0027 - 0.0054
>0.0054
a The assumed meal size is 8 oz (0.227 kg). The ranges of chemical concentrations presented are conservative,
e.g., the 12-meal-per-month levels represent the concentrations associated with 12 to 15.9 meals.
Notes:
1. Consumption limits are based on adult body weight of 70 kg and a cancer slope factor of 0.35 (mg/kg-d"1).
2. None = No consumption recommended.
3. In cases where >16 meals per month are consumed, refer to Equations 3-1 and 3-2, Section 3.2.1.2, for
methods to determine safe consumption limits.
4. The detection limit for chlordane is 1 x 10~3 mg/kg.
5. Instructions for modifying the variables in this table are found in Section 3.3.
6. ARL = Acceptable risk level.
3.3.2 Cancer Potencies and Chronic Reference Doses (q.,*s and RfDs)
Table 3-1 contains the risk values used in the development of the consumption
limit tables shown in Section 4. All of the CSFs and RfDs were obtained from
EPA databases, primarily from IRIS (1999). Preference was given to IRIS values
because these values represent consensus within EPA. When IRIS values were
not available, RfDs from other EPA sources were used (see Section 5).
3.3.3 Consumer Body Weight (BW)
The consumption limit tables in Section 4 are based on fish consumer body
weight of 70 kg (156 Ib), the average body weight of male and female adults in
the U.S. population (U.S. EPA, 1990a).
3-12
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3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
As Equation 3-3 shows, consumption limits are linearly related to body weight.
That is, the higher the body weight assumed for the population of concern, the
higher the consumption limits. EPA's Exposure Factors Handbook (U.S. EPA,
1990a) provides additional specific body weight information that can be used to
adjust the body weight component of Equation 3-3. The values can also be used
to develop a set of multipliers to directly adjust consumption limits for body weight
variations.
Table 3-5 provides a range of average body weights (based on age and sex) for
the U.S. population and their associated multipliers. Values in bold are those
values used in the calculation of the consumption limit tables in Section 4. A
multiplier is provided for each age group, which represents the number by which
the meal consumption limits in the general adult population tables may be
multiplied to calculate new meal consumption limits using an alternative body
weight.
3.3.3.1 Derivation of Multipliers for Body Weight Adjustment—
Body weight multipliers represent the ratio of the alternative body weight to the
standard 70-kg adult body weight. Body weight multipliers were calculated as
follows:
A * i .u • i • Al ter native Consumer Body Weight
Multiplier™, = ^ ^—(3-5)
BW General Adult Body Weight v ;
To derive modified consumption limits using alternative values for body weight,
multiply the existing consumption limits (in meals per month) found in the tables
for the 70-kg adult fisher consumer by the multiplier associated with the new
body weight:
CRmm = CRmm70_kgBW ' MultiplierBW (3.6)
where
Crmm = maximum allowable fish consumption rate (meals/mo)
CRmm = maximum allowable fish consumption rate of a 70-kg
mm70-kgBW ,. . / i / \
fish consumer (meals/mo)
BW = consumer body weight (kg)
MultiplierBW = body weight multiplier (unitless).
3.3.4 Meal Size
Meal size is defined as the amount of fish (in kilograms) consumed at one meal.
EPA has identified a value of 8 oz (227 g) of uncooked fish fillet per 70-kg
consumer body weight as an average meal size for adults in the general
population assuming consumption of noncommercially caught fish only. At this
3-13
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3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
Table 3-5. Average Body Weights and Associated Multipliers
Age Group Average Male
(yr)a Body Weight (kg)
<3
3 to 6
Oto6
6 to 9
9 to 12
12to15
15 to 18
18 to 25
25 to 35
35 to 45
45 to 55
55 to 65
65 to 75
18 to 45
18 to 75
11.9
17.6
14.8
25.3
35.7
50.5
64.9
73.8
78.7
80.9
80.9
78.8
74.8
—
78.1
Average Female
Body Weight (kg)
11.2
17.1
14.2
24.6
36.2
50.7
57.4
60.6
64.2
67.1
68.0
67.9
66.6
64
65.4
Average Body Weight for
Males and Females
Combined (kg)
11.6
17.4
14.5
25.0
36.0
50.6
61.2
67.2
71.5
74.0
74.5
73.4
70.7
—
71.8C(70)C
Multiplier13
0.17
0.25
0.21
0.36
0.51
0.72
0.87
0.96
1.0
1.1
1.1
1.0
1.0
0.91
1.0
a Numbers in bold represent the default values used to calculate the consumption limit tables.
b The body weight multiplier is multiplied by the consumption limits associated with 72-kg adult fish consumers
to obtain new consumption limits using the alternative body weight (see Section 3.3.3). The body weight
multiplier represents the alternative body weight divided by the adult body weight.
c Per recommendations in the Exposure Factors Handbook, the body weight value of 71.8 kg was rounded to 70
kg (U.S. EPA, 1990a).
EPA recommends that the same default value be used for shellfish. However,
EPA is currently investigating this issue and a different default value may be
recommended in the future. Readers may wish to develop fish consumption limits
using other meal sizes obtained from data on local fish consumption patterns
and/or other fish consumption surveys as appropriate (see Appendix B). Table
3-6 provides alternative meal sizes and their associated multipliers. To obtain
modified consumption limits using alternative values for meal size, multiply the
existing consumption limits found in the tables for the 8-oz meal size by the
multiplier associated with the new meal size:
New CR = CR
mm mms
Multiplier
MS
(3-7)
where variables are as previously defined.
3-14
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3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
In addition, if specific meal consumption limits are desired for consumers ages
4 to adult, modifications can be made for both body weight and meal size using
the following equation:
New CR = CR • Multiplier,,,,, • Multiplies..^
mm mm7o kg BW.S-OZ MS F BW F M^ta$
where the parameters are as previously defined.
3.3.5 Contaminant Concentration in Fish Tissue
Chemical contaminant concentrations in fish tissue are influenced by the specific
species and age (size) class of the fish sampled, the chemical properties of the
chemical contaminant (e.g., degradation rate, solubility, bioconcentration poten-
tial), and the contaminant level in the waterbody. A detailed discussion of
selection of target species for use in fish sampling and analysis programs is
presented in Section 3 of Volume 1 of this guidance series. In addition, the
reader may obtain some indication of the range of contaminant concentrations
possible for a specific target analyte in a specific species by reviewing results of
regional and national fish sampling programs such as the EPA National Study of
Chemical Residues in Fish (U.S. EPA, 1991 b), The National Contaminant
Biomonitoring Program (Kidwell et al., 1995), the U.S. Fish and Wildlife Service
National Contaminant Biomonitoring Program (Lowe et al., 1985; Schmitt et al.,
1990), and the National Oceanic and Atmospheric Association (NOAA) Status
and Trends Program (NOAA, 1989).
Note: The chemical contaminant concentration in fish tissue values used in
calculating the risk-based consumption limits should be derived from monitoring
data obtained from fish sampling and analysis programs and be specific to the
waterbody, fish species, and fish size (age) class that were sampled.
3.3.6 Modifying Time-Averaging Period (Tap)
Calculated daily consumption limits represent the maximum amount of fish (in
kilograms) expected to generate a risk no greater than the maximum ARL used
for carcinogens or the maximum amount of fish (in kilograms) that would be
expected not to cause adverse noncarcinogenic health effects based on a
lifetime of daily consumption at that consumption rate. Most fish consumers,
however, do not think about consumption in kilograms per day. Therefore, con-
sumption limits may be more conveniently communicated to the fish-consuming
public expressed as the allowable number of fish meals of a specified meal size
that may be consumed over a given time period.
Monthly consumption limits were derived for all target analytes as shown in
Section 4. For chemical contaminants with carcinogenic properties, there is no
current methodology for evaluating the difference in cancer risks between
consuming a large amount of the carcinogenic contaminant over a short period
of time and consuming the same amount over the course of a lifetime. Therefore,
EPA's current cancer risk assessment guidelines recommend prorating exposure
3-15
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3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
over the lifetime of the exposed individual (U.S. EPA, 1986a). To provide usable
and easily understood consumption guidance, the time-averaging period of 1
month was used as the basis for expressing meal consumption limits in
Section 4. In certain situations, risk managers may wish to calculate alternate
consumption limits for different time intervals. For example, the state of
Minnesota calculates consumption limits for mercury for 3-week (vacation), 3-
month (seasonal), and annual time periods. This is done for mercury because
it is eliminated from the body in a relatively short time period (half-life of
approximately 50 days) and also because of seasonal fish consumption patterns
in the state.
3.4 MODIFICATION OF CONSUMPTION LIMITS FOR A SINGLE CONTAMINANT
IN A MULTISPECIES DIET
Equations 3-1 and 3-3 may be modified to calculate consumption limits for
exposure to a single contaminant through consumption of several different fish
species. This section describes the modifications required to do this.
Individuals often eat several species of fish in their diets. Equations 3-1 and 3-3,
however, are based on contaminant concentrations in a single species of fish.
Where multiple species of contaminated fish are consumed by a single individual,
such limits may not be sufficiently protective. If several fish species are
contaminated with the same chemical, then doses from each of these species
must first be summed across all species eaten in proportion to the amount of
each fish species eaten. This is described by Equation 3-9:
mJ-Pj (3-9)
j=i
where
Ctm = total concentration of chemical contaminant m in an individual's
fish diet (mg/kg)
Cmj = concentration of chemical contaminant m in species; (mg/kg)
Pj = proportion of species; in the diet (unitless).
Note: This equation requires that the risk assessor know or be able to estimate
the proportion of each fish species in the exposed individual's diet. Equation 3-9
yields the weighted average contaminant concentration across all fish species
consumed (Ctm), which then may be used in modified versions of Equations 3-1
to 3-3 to calculate overall and species-specific risk-based consumption limits for
carcinogenic and noncarcinogenic effects as shown in Sections 3.4. 1 and 3.4.2.
3.4.1 Carcinogenic Effects
The equation to calculate an overall daily consumption limit based on exposure
to a single carcinogen in a multiple species diet is very similar to Equation 3-1.
However, in place of Cm, which indicates the average chemical contaminant
concentration in one species, Equation 3-10 uses the equation for Ctm, the
3-16
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3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
weighted average chemical contaminant concentration across all of the species
consumed:
_ ARL « BW
Hm " (3-10)
(cmj • Pj) • CSF (310)
where
CR|im = maximum allowable fish consumption rate (kg/d)
ARL = maximum acceptable lifetime risk level (unitless)
BW = consumer body weight (kg)
Cmj = concentration of chemical contaminant m in fish species; (mg/kg)
Pj = proportion of a given species in the diet (unitless)
CSF = cancer slope factor, usually the upper 95 percent confidence limit
on the linear term in the multistage model used by EPA ([mg/kg-
d])-1).
The daily consumption limit for each species is then calculated as:
CRj =CRllm^Pj (3-11)
where
Cr] = consumption rate of fish species; (kg/d)
CR|im = maximum allowable fish consumption rate (kg/d)
Pj = proportion of a given species in the diet (unitless).
Meal consumption limits may then be calculated for each species as before using
Equation 3-2 (see Section 3.2), with CRj substituted for CR,im in the equation.
Note that Equation 3-11 may be used before or after Equation 3-2, with the same
results.
3.4.2 Noncarcinogenic Effects
For noncarcinogenic effects, the equation to calculate an overall daily
consumption limit based on exposure to a single noncarcinogenic chemical in a
multiple species diet is similar to Equation 3-3 for a single species. However, in
place of Cm, which indicates the chemical contaminant concentration in one
species, Equation 3-12 uses the equation for Ctm, the weighted average chemical
contaminant concentration across all of the species consumed:
RfD « BW
n (3-12)
E (C • • P.) ( }
Z.^ v mj y
3-17
-------
3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
where the parameters are as defined above. The consumption rate for each
species is then calculated using Equation 3-11. Meal consumption limits for each
species may then be calculated as before using Equation 3-2.
3.5 MODIFICATION OF CONSUMPTION LIMITS FOR MULTIPLE CONTAMINANT
EXPOSURES
Equations 3-10 and 3-12 discussed in Section 3.4 can be further modified to
develop consumption limits for multiple chemical exposures across single or
multiple fish species. Section 2.3.4 provides additional information on exposure
to multiple chemical contaminants.
Individuals who ingest chemically contaminated fish may be exposed to a number
of different chemicals simultaneously. This could occur when: (1) a single fish
species is contaminated with several different chemical contaminants; (2) an
individual consumes a mixture of species in his or her diet, each contaminated
with a different chemical; or (3) some combination of the above circumstances
occurs.
EXAMPLE 10:
Calculating Consumption Limits fora Single
Contaminant in a Multispecies Diet
The combined results from a fish sampling and analysis program and a local
fish consumption survey determine that local fishers eat a diet of 30 percent
catfish contaminated with 0.006 mg/kg chlordane and 70 percent trout con-
taminated with 0.008 mg/kg chlordane. The RfD for chlordane reported in IRIS
is 0.00005 mg/kg/d (IRIS, 1999). Because chlordane causes both chronic
health and carcinogenic effects, consumption limits must be calculated for both
health endpoints. The CSF for chlordane reported in IRIS is 0.35 per (mg/kg-
d)"1 (IRIS, 1999). The average body weight of an adult is estimated to be 70
kg.
Carcinogenic Effects: Using a risk level of 10~5 and the values specified
above, Equation 3-5 yields a daily consumption rate of 0.028 kg/d, based on
carcinogenic endpoints:
CR,im
10
-5
70 kg
(0.006 mg/kg • 0.3 + 0.008 mg/kg • 0.7) • 0.35 per mg/kg-d
= 0.029 kg/d .
Equation 3-2 is then used as before to calculate a monthly meal consumption
limit, based on a meal size of 8 oz (0.227 kg):
CR,
0.029 kg-d • 30.44 d/mo
0.227 kg/meal
= 38.8 ~ 39 meals/mo .
(continued)
3-18
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3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
EXAMPLE 10 (continued)
Equation 3-2 yields a meal consumption limit of 39 8-oz meals per month
based on chlordane's carcinogenicity.
Based on a diet of 70 percent trout and 30 percent catfish:
'Rtrout = 39 8-oz meals/mo • 0.7 = 27 8-oz meals/mo
An adult may safely consume 27 8-oz meals of trout and 12 8-oz meals of
catfish per month.
Note: In both cases the meal consumption limits were rounded down. This is
a conservative approach. One might also round up the number of meals of
the species with the lower contaminant concentration, and round down the
number of meals of the species with the higher contaminant concentration, so
that the total number of fish meals per month equals that found by using
Equations 3-6 and 3-2.
Noncarcinogenic Effects: Equation 3-8 is used to calculate the daily
consumption limit based on chlordane's noncarcinogenic health effects using
the RfD rather than the CSF
c
70 kg
0.006 mg/kg • 0.3 + 0.008 mg/kg • 0.7
As with carcinogenic effects, Equation 3-2 is used to convert the daily
consumption limit of 0.570 kg fish to a meal consumption limit:
mm
4.73 kg/d « 30.44 d/mo
0.227 kg/meal
ro. 0 ro, , ,
= 634.3 ~ 634 meals/mo
This analysis indicates that 4.73 kg/d is equivalent to 634 8-oz fish meals per
month or over two 8-oz fish meals per day under this mixed-species diet. This
is categorized as safe fish consumption (represented by ">16" meals/ month)
and has been defined as an intake limit of 16 meals per month for the monthly
consumption limit tables in Section 4. Thus, based on the above results, risk
managers might choose to issue a consumption advisory for adults based on
chlordane's carcinogenic effects, the more sensitive of the two health
endpoints.
Possible toxic interactions in mixtures of chemicals are usually placed in one of
three categories:
3-19
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3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
Antagonistic—the chemical mixture exhibits less toxicity than the chemicals
considered individually
Synergistic—the chemical mixture is more toxic than the sum of the
individual toxicities of the chemicals in the mixture
Additive—the toxicity of the chemical mixture is equal to the sum of the
toxicities of the individual chemicals in the mixture.
Using available data is especially important in cases where mixtures exhibit
synergistic interactions, thereby increasing toxicity. Very little data are available
on the toxic interactions between multiple chemicals, however, and no
quantitative data on interactions between any of the target analytes considered
in this document were located. Some qualitative information is provided in
Section 2.3.4.
If all of the chemicals in a mixture induce the same health effect by similar modes
of action (e.g., cholinesterase inhibition), contaminants may be assumed to
contribute additively to risk (U.S. EPA, 1986c), unless specific data indicate
otherwise. Chemicals in a particular class (e.g., organochlorine or organophos-
phate pesticides) usually have similar mechanisms of toxicity and produce similar
effects. Effects of chemicals and chemical groups are discussed in more detail
in Section 5. For mixtures of chemicals that produce similar toxicological
endpoints, EPA recommends dose addition. This procedure involves scaling the
doses of the components for potency and adding the doses together; the
mixtures response is then estimated for the combined dose (U.S. EPA, 1999a).
Some chemical mixtures may contain chemicals that produce dissimilar health
effects. For these chemicals, EPA recommends response addition. This
procedure involves first determining the risks for the exposure for the individual
components; the mixture risk is then estimated by adding the individual risks
together (U.S. EPA, 1999a).
3.5.1 Carcinogenic Effects
Few empirical studies have considered response addition in any depth, and few
studies have modeled cancer risk from joint exposure. If interactions data are
available on the components of the chemical mixture, EPA recommends that they
be incorporated into the risk assessment by using the interactions-based hazard
index or by including a qualitative assessment of the direction and magnitude of
the impact of the interaction data (U.S. EPA, 1999a).
A detailed discussion of the interactions-based hazard index approach is
available in EPA's proposed guidance for conducting health risk assessment of
chemical mixtures (U.S. EPA, 1999a). For calculating consumption limits,
additivity will be assumed for both carcinogenic and noncarcinogenic effects of
components of chemical mixtures.
Equation 3-13 can be used to calculate a daily consumption rate for chemical
mixtures of carcinogens in single or multiple fish species. It is similar to
3-20
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3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
Equation 3-1, with the summation of all species and all chemicals substituted for
Cm in the denominator:
ARL • BW
(3-13)
• CSF
where
CR|im = maximum allowable fish consumption rate (kg/d)
ARL = maximum acceptable lifetime risk level (unitless)
BW = consumer body weight (kg)
Cmj = concentration of chemical contaminant m in species; (mg/kg)
PJ = proportion of a given species in the diet (unitless)
CSF = cancer slope factor, usually the upper 95 percent confidence limit
on the linear term in the multistage model used by EPA ([mg/
kg-d]-1).
Meal consumption limits for mixtures of carcinogens are then calculated using
Equation 3-2. When only one fish species is involved, Equation 3-13 may be
simplified to Equation 3-14:
= ARL « BW
"" "EC
...
m=l
where the variables are as previously defined.
3.5.2 Noncarcinogenic Effects
Equation 3-15 can be used to calculate a daily consumption rate for noncarcino-
genic chemical mixtures in single or multiple fish species. It is similar to Equation
3-3, with the summation of all species and all chemicals assumed to act
additively. Equation 3-3 has been modified with the respective summation of
concentrations (Cmj) substituted in the denominator and their respective RfDs in
the numerator.
* (RfD 'P \
CR, = Y, • BW ,3 15x
i \ (C • P M (J-1£v
™=1 ^ <^mj *V )
where the parameters are as previously defined and Pm = proportion by weight
of chemical in diet. Meal consumption limits are then calculated using Equation
3-2, as above. Again, when only one fish species is involved, Equation 3-15 can
be simplified to Equation 3-16:
3-21
-------
3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
CRllm = E
Rf D • P
m m
m=l , -m
c
• BW (3-16)
where the variables are as previously defined. Note that Equations 3-15 and 3-16
may not be used for contaminants causing dissimilar noncarcinogenic health
effects.
EXAMPLE 11: Calculating Consumption Limits for Multiple
Contaminants in a Single Species Diet
A single fish species is contaminated with 0.04 mg/kg chlordane and 0.01
mg/kg heptachlor epoxide. A maximum acceptable risk level of 10~5 and an
adult body weight of 72 kg are used. Because chlordane and heptachlor
epoxide cause both carcinogenic and chronic systemic health effects, both
health endpoints must be considered in establishing consumption limits for
these chemicals.
Carcinogenic Effects: The CSF for chlordane reported in IRIS is 0.35 per
(mg/kg-d) (IRIS, 1999). The CSF for heptachlor epoxide reported in IRIS is 9.1
per (mg/kg-d) (IRIS, 1999). Equation 3-10 is used to calculate daily
consumption rate based on the combined carcinogenic effects of both
contaminants:
IfT5 • 70
CR,. = — — = 0.007 kg/d .
lim (0.04 • 0.35) + (0.01 • 9.1)
A daily consumption rate of 0.007 kg fish per day is calculated. Using
Equation 3-2, this daily consumption rate is converted to a meal
consumption limit of one 4-oz meal per month (or six 8-oz meals per year).
Noncarcinogenic Effects: Chlordane and heptachlorare both organochlorine
pesticides and cause many similar noncarcinogenic effects. Heptachlor
epoxide is a metabolite of the organochlorine pesticide, heptachlor. When
heptachlor is released into the environment, it quickly breaks down into
heptachlor epoxide. Therefore, the toxicity values used in this document are
for heptachlor epoxide, not heptachlor (see Section 5.3.7). Adverse liver
effects formed the basis of the RfDs for both chemicals (IRIS, 1999). A
combined daily consumption limit based on an RfD of 5 x 10~4 mg/kg-d for
chlordane and 1.3 x 1 o~5 mg/kg-d for heptachlor was calculated using Equation
3-12:
(Continued)
3-22
-------
3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
EXAMPLE 11 (continued)
5 x 10"
mg/kg-d 1.3 x 10 5 mg/kg-d
0.04 mg/kg
0.01 mg/kg
• 70 kg = 0.97 kg/d
Equation 3-12 yields a daily consumption rate of 0.97 kg fish/d at the con-
taminant concentrations described above. Using Equation 3-2, a meal
consumption limit of 130 4-oz meals per month is calculated. Therefore, based
on the carcinogenic and chronic systemic consumption limits calculated for
combined heptachlor epoxide and chlordane contamination, a risk manager
may choose to advise (1) limiting fish consumption to six 8-oz meals per year,
based on the combined carcinogenic effects; or (2) limiting fish consumption
to 1334-oz-meals/month, based on noncarcinogenic effects. In general, EPA
advises that the more protective meal consumption limit (in this case, the
limit for the carcinogenic effect) serve as the basis for a fish consumption
advisory to be protective of both health effects endpoints.
3.5.3 Species-Specific Consumption Limits in a Multiple Species Diet
Equation 3-11 is used to calculate the risk-based consumption limits for each
species in a multiple species diet, for both carcinogenic and noncarcinogenic
toxicity where the variables are as defined above. CR,im is calculated using
Equations 3-13 or 3-15, for carcinogenic and noncarcinogenic toxicity,
respectively. As with the consumption limits for single chemicals, these con-
sumption limits are valid only if the assumed mix of species in the diet is known
and if the contaminant concentrations in each species are accurate.
EXAMPLE 12:
Calculating Consumption Limits for Multiple
Contaminants in a Multispecies Diet
Chlorpyrifos and diazinon both cause cholinesterase inhibition, so are con-
sidered together when developing meal consumption limits. The RfD for chlor-
pyrifos is 0.0003 mg/kg-d, (EPA, 2000b), and the RfD for diazinon is 0.0007
mg/kg/d (U.S. EPA, 1998b).
A local fish consumption survey reveals that adult fishers consume trout and
catfish at a ratio of 70:30, respectively. A fish sampling and analysis program
reports chlorpyrifos and diazinon contamination in both species. Trout fillets
are contaminated with 4.0 mg/kg chlorpyrifos and 0.3 mg/kg diazinon. Catfish
fillets are contaminated with 6.0 mg/kg chlorpyrifos and 0.8 mg/kg diazinon.
Given an adult body weight of 70 kg, a risk-based consumption rate of 0.15 kg
fish per day is calculated using Equation 3-11:
(Continued)
3-23
-------
3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
EXAMPLE 12 (continued)
( 0.0003 0.0007
llm (4.0 • 0.7) + (6.0 • 0.3) (0.3 • 0.7) + (0.8 • 0.3)
= 0.11 kg/d .
Using Equation 3-2, a meal consumption limit of 15 8-oz meals per month is
derived. Note: If chlorpyrifos and diazinon did not cause the same health
endpoint, then separate meal consumption limits would have to be calculated
for each as described in Section 3.4.2, with the more protective meal
consumption limit usually serving as the basis for a fish consumption advisory
(see Section 3.5.2).
Based on a diet of 70 percent trout and 30 percent catfish:
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
SECTION 4
RISK-BASED CONSUMPTION LIMIT TABLES
4.1 OVERVIEW AND SECTION ORGANIZATION
This section provides consumption limit tables for carcinogenic and chronic health
endpoints for the general adult population for all of the target analytes listed in
Table 1-1.
Variables used to calculate the consumption limits include fish meal size,
consumer body weight, contaminant concentration in the fish tissue, the time-
averaging period selected (monthly), the reference dose for noncarcinogenic
health endpoints, and the cancer potency factor and the maximum acceptable risk
level for carcinogenic health endpoints. Default values for the variables are
presented in Section 3 and described in greater detail in Section 2.
Each consumption table lists, by chemical, the maximum number of fish meals per
unit time (monthly) that may be safely eaten. Readers may use these tables by:
determining the chemical contaminant concentration in fish surveyed in local fish
sampling and analysis programs and reading the value for the maximum number
of meals per month that may be safely eaten for each contaminant for noncancer
and cancer endpoints. For those contaminants with monthly fish consumption
limits calculated for both the noncancer and cancer endpoints, EPA recommends
using the more conservative of the two values. In cases where >16 meals per
month are consumed, refer to Equations 3-1 and 3-2, Section 3.2.1.2, for methods
to determine safe consumption limits.
Some of the contaminant concentrations shown in the consumption limit tables
are below current laboratory detection limits. Because of improvements in
chemical analysis procedures and associated technologies, however, chemical
detection limits regularly decrease. The fish tissue concentrations that are
currently below the limit of detection are provided so that risk managers may use
them once lower detection limits are achievable through improvements in
analytical procedures. Note: The reader should be aware that detection limits
presented here are derived from state-of-the-art state, regional, and national fish
monitoring programs and may not be representative of detection limits achievable
in all laboratories. Readers should consult with the analytical chemists in their
state responsible for analyzing fish tissue samples to ensure that their detection
limits are comparable to those presented. If the detection limits presented are
lower than those achieved in the state's program, the reader should make
4-1
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
necessary adjustments to the tables. The detection limits presented here are to
provide general guidance on detection limits typically achievable using current
analytical procedures. The reader should review Section 6 of Volume 1 for further
information on chemical analysis procedures and associated detection and
quantitation limits for the target analytes.
4.2 CONSUMPTION LIMIT TABLES
Tables 4-1 through 4-25 are consumption limit tables for carcinogenic and chronic
systemic health endpoints for each of the target analytes. Readers using the
tables as a basis for fish consumption advisories should note that the values given
in the tables are valid only for single contaminants in single-species diets.
Sections 3.4 and 3.5 describe methods for calculating consumption limits for
multiple contaminant situations and for multiple fish species diets.
4-2
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-1. Monthly Fish Consumption Limits for Carcinogenic and Noncarcinogenic
Health Endpoints - Arsenic (inorganic)
Risk Based Consumption
Limit3
Fish Meals/Month
Unrestricted (>16)
16
12
8
4
3
2
1
0.5
None (<0.5)
Noncancer Health Endpointsb
Fish Tissue Concentrations,
(ppm, wet weight)
0-0.088
>0.088-0.18
>0.18-0.23
>0.23-0.35
>0.35-0.7
>0.7-0.94
>0.94-1.4
>1.4-2.8
>2.8-5.6
>5.6
Cancer Health Endpoints0
Fish Tissue Concentrations
(ppm, wet weight)
0 - 0.002
>0.002 - 0.0039
>0.0039 - 0.0052
>0.0052 - 0.0078
>0.0078- 0.016
>0.016- 0.021
>0.021 -0.031
>0.031 - 0.063
>0.063-0.13
>0.13
a The assumed meal size is 8 oz (0.227 kg). The ranges of chemical concentrations presented are conservative, e.g., the
12-meal-per-month levels represent the concentrations associated with 12 to 15.9 meals.
b Chronic, systemic effects.
c Cancer values represent tissue concentrations at a 1 in 100,000 risk level.
Notes:
1. Consumption limits are based on an adult body weight of 70 kg, an RfD of 3x10"4 mg/kg-d, and a cancer slope factor
(CSF)of 1.5(mg/kg-d)-1
2. None = No consumption recommended.
3. In cases where >16 meals per month are consumed, refer to Equations 3-1 and 3-2, Section 3.2.1.2, for methods to
determine safe consumption limits.
4. The detection limit for arsenic is 5 x 10"3 mg/kg.
5. Instructions for modifying the variables in this table are found in Section 3.3.
6. Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD). When the monthly limit
is consumed in less than 1 month (e.g., in a few large meals), the daily dose may exceed the RfD (see Section 2.3).
4-3
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-2. Monthly Fish Consumption Limits for Noncarcinogenic Health
Endpoint - Cadmium
Risk Based Consumption Limit3
Fish Meals/Month
Unrestricted (>16)
16
12
8
4
3
2
1
0.5
None (<0.5)
Noncancer Health Endpointsb
Fish Tissue Concentrations
(ppm, wet weight)
0-0.088
>0.088-0.18
>0. 18 -0.23
>0.23-0.35
>0.35-0.7
>0.7-0.94
>0.94-1.4
>1.4-2.8
>2.8-5.6
>5.6
a The assumed meal size is 8 oz (0.227 kg). The ranges of chemical concentrations presented are conservative, e.g., the
12-meal-per-month levels represent the concentrations associated with 12 to 15.9 meals.
b Chronic, systemic effects.
Notes:
1. Consumption limits are based on an adult body weight of 70 kg and an RfD of 1x10"3 mg/kg-d.
2. None = No consumption recommended.
3. In cases where >16 meals per month are consumed, refer to Equations 3-1 and 3-2, Section 3.2.1.2, for methods to
determine safe consumption limits.
4. The detection limit for cadmium is 5 x 10"3 mg/kg.
5. Instructions for modifying the variables in this table are found in Section 3.3.
6. Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD). When the monthly limit
is consumed in less than 1 month (e.g., in a few large meals), the daily dose may exceed the RfD (see Section 2.3).
4-4
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-3. Monthly Fish Consumption Limits for Noncarcinogenic Health Endpoint -
Methylmercury
Risk Based Consumption Limit3
Fish Meals/Month
Unrestricted (>16)
16
12
8
4
3
2
1
0.5
None (<0.5)
Noncancer Health Endpointsb
Fish Tissue Concentrations
(ppm, wet weight)
0 - 0.029
>0. 029 -0.059
>0.059- 0.078
>0.078-0.12
>0. 12 -0.23
>0.23-0.31
>0.31 -0.47
>0.47-0.94
>0.94- 1.9
>1.9
a The assumed meal size is 8 oz (0.227 kg). The ranges of chemical concentrations presented are conservative, e.g., the
12-meal-per-month levels represent the concentrations associated with 12 to 15.9 meals.
b Chronic, systemic effects.
Notes:
1. Consumption limits are based on an adult body weight of 70 kg and an interim RfD of 1x10"4 mg/kg-d.
2. None = No consumption recommended.
3. In cases where >16 meals per month are consumed, refer to Equations 3-1 and 3-2, Section 3.2.1.2, for methods to
determine safe consumption limits.
4. The detection limit for methylmercury is 1 x 10"3 mg/kg.
5. Instructions for modifying the variables in this table are found in Section 3.3.
6. Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD). When the monthly limit
is consumed in less than 1 month (e.g., in a few large meals), the daily dose may exceed the RfD (see Section 2.3).
4-5
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-4. Monthly Fish Consumption Limits for Noncarcinogenic Health Endpoint -
Selenium
Risk Based Consumption Limit3
Fish Meals/Month
Unrestricted (>16)
16
12
8
4
3
2
1
0.5
None (<0.5)
Noncancer Health Endpointsb
Fish Tissue Concentrations
(ppm, wet weight)
0-1.5
>1.5-2.9
>2.9-3.9
>3.9-5.9
>5.9-12
>12-16
>16-23
>23 - 47
>47 - 94
>94
a The assumed meal size is 8 oz (0.227 kg). The ranges of chemical concentrations presented are conservative,
e.g., the 12-meal-per-month levels represent the concentrations associated with 12 to 15.9 meals.
b Chronic, systemic effects.
Notes:
1. Consumption limits are based on an adult body weight of 70 kg and an RfD of 5x10"3 mg/kg-d.
2. None = No consumption recommended.
3. In cases where >16 meals per month are consumed, refer to Equations 3-1 and 3-2, Section 3.2.1.2, for methods
to determine safe consumption limits.
4. The detection limit for selenium is 17x10"3 mg/kg.
5. Instructions for modifying the variables in this table are found in Section 3.3.
6. Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD). When the
monthly limit is consumed in less than 1 month (e.g., in a few large meals), the daily dose may exceed the RfD
(see Section 2.3).
4-6
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-5. Monthly Fish Consumption Limits for Noncarcinogenic Health Endpoint
Tributyltin
Risk Based Consumption Limit3
Fish Meals/Month
Unrestricted (>16)
16
12
8
4
3
2
1
0.5
None (<0.5)
Noncancer Health Endpointsb
Fish Tissue Concentrations
(ppm, wet weight)
0-0.088
>0. 088 -0.18
>0.18-0.23
>0.23-0.35
>0.35-0.7
>0.7-0.94
>0.94-1.4
>1.4-2.8
>2.8-5.6
>5.6
a The assumed meal size is 8 oz (0.227 kg). The ranges of chemical concentrations presented are conservative,
e.g., the 12-meal-per-month levels represent the concentrations associated with 12 to 15.9 meals.
b Chronic, systemic effects.
Notes:
1. Consumption limits are based on an adult body weight of 70 kg and an RfD of 3x10"4 mg/kg-d.
2. None = No consumption recommended.
3. In cases where >16 meals per month are consumed, refer to Equations 3-1 and 3-2, Section 3.2.1.2, for methods
to determine safe consumption limits.
4. The detection limit for tributyltin is 2x10"3 mg/kg.
5. Instructions for modifying the variables in this table are found in Section 3.3.
6. Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD). When the
monthly limit is consumed in less than 1 month (e.g., in a few large meals), the daily dose may exceed the RfD
(see Section 2.3).
4-7
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-6. Monthly Fish Consumption Limits for Carcinogenic and Noncarcinogenic
Health Endpoints - Chlordane
Risk Based Consumption Limit3
Fish Meals/Month
Unrestricted (>16)
16
12
8
4
3
2
1
0.5
None (<0.5)
Noncancer Health Endpointsb
Fish Tissue Concentrations
(ppm, wet weight)
0-0.15
>0. 15 -0.29
>0.29-0.39
>0. 39 -0.59
>0.59-1.2
>1.2- 1.6
>1.6-2.3
>2.3-4.7
>4.7-9.4
>9.4
Cancer Health Endpoints0
Fish Tissue Concentrations
(ppm, wet weight)
0 - 0.0084
>0. 0084 -0.01 7
>0.017-0.022
>0.022 - 0.034
>0.034 - 0.067
>0. 067 -0.089
>0.089-0.13
>0. 13 -0.27
>0.27-0.54
>0.54
a The assumed meal size is 8 oz (0.227 kg). The ranges of chemical concentrations presented are conservative, e.g., the
12-meal-per-month levels represent the concentrations associated with 12 to 15.9 meals.
b Chronic, systemic effects.
c Cancer values represent tissue concentrations at a 1 in 100,000 risk level.
Notes:
1. Consumption limits are based on an adult body weight of 70 kg, an RfD of 5x10"4 mg/kg-d, and a cancer slope factor
(CSF)of0.35(mg/kg-d)-1
2. None = No consumption recommended.
3. In cases where >16 meals per month are consumed, refer to Equations 3-1 and 3-2, Section 3.2.1.2, for methods to
determine safe consumption limits.
4. The detection limit for chlordane is 1 x 10"3 mg/kg.
5. Instructions for modifying the variables in this table are found in Section 3.3.
6. Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD). When the monthly limit
is consumed in less than 1 month (e.g., in a few large meals), the daily dose may exceed the RfD (see Section 2.3).
4-8
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-7. Monthly Fish Consumption Limits for Carcinogenic and Noncarcinogenic
Health Endpoints - DDT
Risk Based Consumption Limit3
Fish Meals/Month
Unrestricted (>16)
16
12
8
4
3
2
1
0.5
None (<0.5)
Noncancer Health Endpointsb
Fish Tissue Concentrations
(ppm, wet weight)
0-0.015
>0. 015 -0.029
>0.029 - 0.039
>0.039- 0.059
>0.059-0.12
>0. 12 -0.16
>0.16-0.23
>0.23 - 0.47
>0.47 - 0.94
>0.94
Cancer Health Endpoints0
Fish Tissue Concentrations
(ppm, wet weight)
0 - 0.0086
>0.0086- 0.017
>0.017- 0.023
>0.023 - 0.035
>0.035 - 0.069
>0.069 - 0.092
>0.092-0.14
>0. 14 -0.28
>0.28-0.55
>0.55
a The assumed meal size is 8 oz (0.227 kg). The ranges of chemical concentrations presented are conservative, e.g.,
the 12-meal-per-month levels represent the concentrations associated with 12 to 15.9 meals.
b Chronic, systemic effects.
c Cancer values represent tissue concentrations at a 1 in 100,000 risk level.
Notes:
1. Consumption limits are based on an adult body weight of 70 kg, an RfD of 5x10"4 mg/kg-d, and a cancer slope factor
(CSF)of0.34(mg/kg-d)-1
2. None = No consumption recommended.
3. In cases where >16 meals per month are consumed, refer to Equations 3-1 and 3-2, Section 3.2.1.2, for methods to
determine safe consumption limits.
4. The detection limit for DDT is 1 x 10"4 mg/kg.
5. Instructions for modifying the variables in this table are found in Section 3.3.
6. Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD). When the monthly limit
is consumed in less than 1 month (e.g., in a few large meals), the daily dose may exceed the RfD (see Section 2.3).
4-9
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4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-8. Monthly Fish Consumption Limits for Noncarcinogenic Health Endpoint
Dicofol
Risk Based Consumption Limit3
Fish Meals/Month
Unrestricted (>16)
16
12
8
4
3
2
1
0.5
None (<0.5)
Noncancer Health Endpointsb
Fish Tissue Concentrations
(ppm, wet weight)
0-0.088
>0.088-0.18
>0. 18 -0.23
>0.23-0.35
>0.35-0.7
>0.7-0.94
>0.94-1.4
>1.4-2.8
>2.8-5.6
>5.6
a The assumed meal size is 8 oz (0.227 kg). The ranges of chemical concentrations presented are conservative,
e.g., the 12-meal-per-month levels represent the concentrations associated with 12 to 15.9 meals.
b Chronic, systemic effects.
Notes:
1. Consumption limits are based on an adult body weight of 70 kg and an RfD of 4x10"4 mg/kg-d.
2. None = No consumption recommended.
3. In cases where >16 meals per month are consumed, refer to Equations 3-1 and 3-2, Section 3.2.1.2, for methods
to determine safe consumption limits.
4. The detection limit for dicofol is 1x10"3 mg/kg.
5. Instructions for modifying the variables in this table are found in Section 3.3.
6. Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD). When the
monthly limit is consumed in less than 1 month (e.g., in a few large meals), the daily dose may exceed the RfD
(see Section 2.3).
4-10
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4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-9. Monthly Fish Consumption Limits for Carcinogenic and Noncarcinogenic
Health Endpoints - Dieldrin
Risk Based Consumption Limit3
Fish Meals/Month
Unrestricted (>16)
16
12
8
4
3
2
1
0.5
None (<0.5)
Noncancer Health Endpointsb
Fish Tissue Concentration
(ppm, wet weight)
0-0.015
>0.015- 0.029
>0.029 - 0.039
>0.039- 0.059
>0.059-0.12
>0. 12 -0.16
>0.16-0.23
>0.23 - 0.47
>0.47 - 0.94
>0.94
Cancer Health Endpoints0
Fish Tissue Concentrations
(ppm, wet weight)
0-0.00018
>0. 0001 8 -0.00037
>0. 00037 - 0.00049
>0.00049 - 0.00073
>0.00073- 0.0015
>0. 0015 -0.002
>0.002 - 0.0029
>0.0029 - 0.0059
>0. 0059 -0.012
>0.012
a The assumed meal size is 8 oz (0.227 kg). The ranges of chemical concentrations presented are conservative, e.g., the
12-meal-per-month levels represent the concentrations associated with 12 to 15.9 meals.
b Chronic, systemic effects.
c Cancer values represent tissue concentrations at a 1 in 100,000 risk level.
Notes:
1. Consumption limits are based on an adult body weight of 70 kg, an RfD of 5x10"5 mg/kg-d, and a cancer slope factor
(CSF)of 16(mg/kg-d)-1
2. None = No consumption recommended.
3. In cases where >16 meals per month are consumed, refer to Equations 3-1 and 3-2, Section 3.2.1.2, for methods to
determine safe consumption limits.
4. The detection limit for dieldrin is 1x10"4 mg/kg.
5. Instructions for modifying the variables in this table are found in Section 3.3.
6. Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD). When the monthly limit
is consumed in less than 1 month (e.g., in a few large meals), the daily dose may exceed the RfD (see Section 2.3).
4-11
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4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-10. Monthly Fish Consumption Limits for Noncarcinogenic Health Endpoint
Endosulfan
Risk Based Consumption Limit3
Fish Meals/Month
Unrestricted (>16)
16
12
8
4
3
2
1
0.5
None (<0.5)
Noncancer Health Endpointsb
Fish Tissue Concentrations
(ppm, wet weight)
0-1.8
>1.8-3.5
>3.5-4.7
>4.7 - 7
>7-14
>14-19
>19-28
>28 - 56
>56-110
>110
a The assumed meal size is 8 oz (0.227 kg). The ranges of chemical concentrations presented are conservative, e.g., the
12-meal-per-month levels represent the concentrations associated with 12 to 15.9 meals.
b Chronic, systemic effects.
Notes:
1. Consumption limits are based on an adult body weight of 70 kg and an RfD of 6x10"3 mg/kg-d.
2. None = No consumption recommended.
3. In cases where >16 meals per month are consumed, refer to Equations 3-1 and 3-2, Section 3.2.1.2, for methods to
determine safe consumption limits.
4. The detection limit for endosulfan is 5 x 10"3 mg/kg.
5. Instructions for modifying the variables in this table are found in Section 3.3.
6. Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD). When the monthly limit
is consumed in less than 1 month (e.g., in a few large meals), the daily dose may exceed the RfD (see Section 2.3).
4-12
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4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-11. Monthly Fish Consumption Limits for Noncarcinogenic Health Endpoint -
Endrin
Risk Based Consumption Limit3
Fish Meals/Month
Unrestricted (>16)
16
12
8
4
3
2
1
0.5
None (<0.5)
Noncancer Health Endpointsb
Fish Tissue Concentrations
(ppm, wet weight)
0-0.088
>0. 088 -0.18
>0.18-0.23
>0.23-0.35
>0.35-0.7
>0.7-0.94
>0.94-1.4
>1.4-2.8
>2.8-5.6
>5.6
a The assumed meal size is 8 oz (0.227 kg). The ranges of chemical concentrations presented are conservative, e.g.,
the 12-meal-per-month levels represent the concentrations associated with 12 to 15.9 meals.
b Chronic, systemic effects.
c Cancer values represent tissue concentrations at a 1 in 100,000 risk level.
Notes:
1. Consumption limits are based on an adult body weight of 70 kg and an RfD of 3x10"4 mg/kg-d.
2. None = No consumption recommended.
3. In cases where >16 meals per month are consumed, refer to Equations 3-1 and 3-2, Section 3.2.1.2, for methods to
determine safe consumption limits.
4. The detection limit for endrin is 1 x 10"4 mg/kg.
5. Instructions for modifying the variables in this table are found in Section 3.3.
6. Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD). When the monthly
limit is consumed in less than 1 month (e.g., in a few large meals), the daily dose may exceed the RfD (see Section
2.3).
4-13
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-12. Monthly Fish Consumption Limits for Carcinogenic and Noncarcinogenic
Health Endpoints - Heptachlor Epoxide
Risk Based Consumption Limit3
Fish Meals/Month
Unrestricted (>16)
16
12
8
4
3
2
1
0.5
None (<0.5)
Noncancer Health Endpointsb
Fish Tissue Concentrations
(ppm, wet weight)
0 - 0.0038
>0.0038 - 0.0076
>0. 0076 -0.01
>0.01 -0.015
>0.015- 0.031
>0.031 -0.041
>0.041 - 0.061
>0.061 -0.12
>0. 12 -0.24
>0.24
Cancer Health Endpoints0
Fish Tissue Concentrations
(ppm, wet weight)
0 - 0.00032
>0. 00032 - 0.00064
>0. 00064 - 0.00086
>0.00086- 0.0013
>0.001 3 -0.0026
>0.0026 - 0.0034
>0.0034 - 0.0052
>0.0052-0.01
>0.01 -0.021
>0.021
a The assumed meal size is 8 oz (0.227 kg). The ranges of chemical concentrations presented are conservative, e.g., the
12-meal-per-month levels represent the concentrations associated with 12 to 15.9 meals.
b Chronic, systemic effects.
c Cancer values represent tissue concentrations at a 1 in 100,000 risk level.
Notes:
1. Consumption limits are based on an adult body weight of 70 kg, an RfD of 1.3x10"5 mg/kg-d, and a cancer slope factor
(CSF)of9.1 (mg/kg-d)-1
2. None = No consumption recommended.
3. In cases where >16 meals per month are consumed, refer to Equations 3-1 and 3-2, Section 3.2.1.2, for methods to
determine safe consumption limits.
4. The detection limit for heptachlor epoxide is 1x10'4 mg/kg.
5. Instructions for modifying the variables in this table are found in Section 3.3.
6. Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD). When the monthly limit
is consumed in less than 1 month (e.g., in a few large meals), the daily dose may exceed the RfD (see Section 2.3).
4-14
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-13. Monthly Fish Consumption Limits for Carcinogenic and Noncarcinogenic
Health Endpoints - Hexachlorobenzene
Risk Based Consumption Limit3
Fish Meals/Month
Unrestricted (>16)
16
12
8
4
3
2
1
0.5
None (<0.5)
Noncancer Health Endpointsb
Fish Tissue Concentrations
(ppm, wet weight)
0-0.23
>0.23 - 0.47
>0.47 - 0.63
>0.63 - 0.94
>0.94-1.9
>1.9-2.5
>2.5-3.8
>3.8-7.5
>7.5-15
>15
Cancer Health Endpoints0
Fish Tissue Concentrations
(ppm, wet weight)
0-0.0018
>0.001 8 -0.0037
>0.0037 - 0.0049
>0.0049 - 0.0073
>0.0073-0.015
>0.015-0.02
>0.02 - 0.029
>0. 029 -0.059
>0.059-0.12
>0.12
a The assumed meal size is 8 oz (0.227 kg). The ranges of chemical concentrations presented are conservative, e.g., the
12-meal-per-month levels represent the concentrations associated with 12 to 15.9 meals.
b Chronic, systemic effects.
c Cancer values represent tissue concentrations at a 1 in 100,000 risk level.
Note:
1. Consumption limits are based on an adult body weight of 70 kg, an RfD of 8x10"4 mg/kg-d, and a cancer slope factor
(CSF) of 1.6 (mg/kg-d)'1
2. None = No consumption recommended.
3. In cases where >16 meals per month are consumed, refer to Equations 3-1 and 3-2, Section 3.2.1.2, for methods to
determine safe consumption limits.
4. The detection limit for hexachlorobenzene is 1 x 10'4 mg/kg.
5. Instructions for modifying the variables in this table are found in Section 3.3.
6. Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD). When the monthly limit
is consumed in less than 1 month (e.g., in a few large meals), the daily dose may exceed the RfD (see Section 2.3).
4-15
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-14. Monthly Fish Consumption Limits for Carcinogenic and Noncarcinogenic
Health Endpoints - Lindane
Risk Based Consumption Limit3
Fish Meals/Month
Unrestricted (>16)
16
12
8
4
3
2
1
0.5
None (<0.5)
Noncancer Health Endpointsb
Fish Tissue Concentrations
(ppm, wet weight)
0-0.088
>0. 088 -0.18
>0. 18 -0.23
>0.23-0.35
>0.35-0.7
>0.7-0.94
>0.94-1.4
>1.4-2.8
>2.8-5.6
>5.6
Cancer Health Endpoints0
Fish Tissue Concentrations
(ppm, wet weight)
0 - 0.0023
>0.0023 - 0.0045
>0.0045 - 0.006
>0.006- 0.009
>0.009- 0.018
>0. 01 8 -0.024
>0.024 - 0.036
>0.036 - 0.072
>0.072-0.14
>0.14
a The assumed meal size is 8 oz (0.227 kg). A range of chemical concentrations are presented that are conservative, e.g.
the 12 meal per month levels represent the concentrations associated with 12 meals up to 15.9 meals.
b Chronic, systemic effects.
c Cancer values represent tissue concentrations at a 1 in 100,000 risk level.
Notes:
1. Consumption limits are based on an adult body weight of 70 kg, an RfD of 3x10"4 mg/kg-d, and a cancer slope factor
(CSF)of 1.3(mg/kg-d)-1
2. None = No consumption recommended.
3. In cases where >16 meals per month are consumed, refer to Equations 3-1 and 3-2, Section 3.2.1.2, for methods to
determine safe consumption limits.
4. The detection limit for lindane is 1 x 10"4 mg/kg.
5. Instructions for modifying the variables in this table are found in Section 3.3.
6. Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD). When the monthly limit
is consumed in less than 1 month (e.g., in a few large meals), the daily dose may exceed the RfD (see Section 2.3).
4-16
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-15. Monthly Fish Consumption Limits for Noncarcinogenic Health Endpoint
Mi rex
Risk Based Consumption Limit3
Fish Meals/Month
Unrestricted (>16)
16
12
8
4
3
2
1
0.5
None (<0.5)
Noncancer Health Endpointsb
Fish Tissue Concentrations
(ppm, wet weight)
0-0.059
>0.059-0.12
>0. 12 -0.16
>0. 16 -0.23
>0.23-0.47
>0.47-0.63
>0.63-0.94
>0.94- 1.9
>1.9-3.8
>3.8
a The assumed meal size is 8 oz (0.227 kg). The ranges of chemical concentrations presented are conservative,
e.g., the 12-meal-per-month levels represent the concentrations associated with 12 to 15.9 meals.
b Chronic, systemic effects.
c Cancer values represent tissue concentrations at a 1 in 100,000 risk level.
Notes:
1. Consumption limits are based on an adult body weight of 70 kg and RfD of 2 x 10"4 mg/kg-d
2. None = No consumption recommended.
3. In cases where >16 meals per month are consumed, refer to Equations 3-1 and 3-2, Section 3.2.1.2, for methods
to determine safe consumption limits.
4. The detection limit for mirex is 1 x 10"4 mg/kg.
5. Instructions for modifying the variables in this table are found in Section 3.3.
6. Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD). When the
monthly limit is consumed in less than 1 month (e.g., in a few large meals), the daily dose may exceed the RfD
(see Section 2.3).
4-17
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-16. Monthly Fish Consumption Limits for Carcinogenic and Noncarcinogenic
Health Endpoints - Toxaphene
Risk Based Consumption Limit3
Fish Meals/Month
Unrestricted (>16)
16
12
8
4
3
2
1
0.5
None (<0.5)
Noncancer Health Endpointsb
Fish Tissue Concentrations
(ppm, wet weight)
0-0.073
>0. 073 -0.15
>0.15-0.2
>0.2 - 0.29
>0.29-0.59
>0.59-0.78
>0.78-1.2
>1.2-2.3
>2.3-4.7
>4.7
Cancer Health Endpoints0
Fish Tissue Concentrations
(ppm, wet weight)
0 - 0.0027
>0.0027 - 0.0053
>0.0053- 0.0071
>0.0071 -0.011
>0.011 -0.021
>0.021 -0.028
>0.028 - 0.043
>0.043- 0.085
>0.085-0.17
>0.17
a The assumed meal size is 8 oz (0.227 kg). The ranges of chemical concentrations presented are conservative, e.g., the
12-meal-per-month levels represent the concentrations associated with 12 to 15.9 meals.
b Chronic, systemic effects.
c Cancer values represent tissue concentrations at a 1 in 100,000 risk level.
Notes:
1. Consumption limits are based on an adult body weight of 70 kg, an RfD of 2.5 x 10"4 mg/kg-d, and a cancer slope factor
(CSF)of 1.1 (mg/kg-d)'1
2. None = No consumption recommended.
3. In cases where >16 meals per month are consumed, refer to Equations 3-1 and 3-2, Section 3.2.1.2, for methods to
determine safe consumption limits.
4. The detection limit for toxaphene is 3 x 10'3 mg/kg.
5. Instructions for modifying the variables in this table are found in Section 3.3.
6. Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD). When the monthly limit
is consumed in less than 1 month (e.g., in a few large meals), the daily dose may exceed the RfD (see Section 2.3).
4-18
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-17. Monthly Fish Consumption Limits for Noncarcinogenic Health Endpoint
Chlorpyrifos
Risk Based Consumption Limit3
Fish Meals/Month
Unrestricted (>16)
16
12
8
4
3
2
1
0.5
None (<0.5)
Noncancer Health Endpointsb
Fish Tissue Concentrations
(ppm, wet weight)
0-0.088
>0.088-0.18
>0. 18 -0.23
>0.23-0.35
>0.35-0.7
>0.7-0.94
>0.94-1.4
>1.4-2.8
>2.8-5.6
>5.6
a The assumed meal size is 8 oz (0.227 kg). The ranges of chemical concentrations presented are conservative,
e.g., the 12-meal-per-month levels represent the concentrations associated with 12 to 15.9 meals.
b Chronic, systemic effects.
Notes:
1. Consumption limits are based on an adult body weight of 70 kg and an RfD of 3x10"4 mg/kg-d.*
2. None = No consumption recommended.
3. In cases where >16 meals per month are consumed, refer to Equations 3-1 and 3-2, Section 3.2.1.2, for methods
to determine safe consumption limits.
4. The detection limit for chlorpyrifos is 2 x 10"3 mg/kg.
5. Instructions for modifying the variables in this table are found in Section 3.3.
6. Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD). When the
monthly limit is consumed in less than 1 month (e.g., in a few large meals), the daily dose may exceed the RfD
(see Section 2.3).
* Because of the potential for adverse neurological development effects, EPA recommends the use of a Population
Adjusted Dose (PAD) of 3x10"5 mg/kg-d for infants, children to age six, and women aged 13-50.
4-19
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-18. Monthly Fish Consumption Limits for Noncarcinogenic Health Endpoint
Diazinon
Risk Based Consumption Limit3
Fish Meals/Month
Unrestricted (>16)
16
12
8
4
3
2
1
0.5
None (<0.5)
Noncancer Health Endpointsb
Fish Tissue Concentrations
(ppm, wet weight)
0-0.21
>0.21 -0.41
>0.41 -0.55
>0.55-0.82
>0.82-1.6
>1.6-2.2
>2.2-3.3
>3.3-6.6
>6.6-13
>13
a The assumed meal size is 8 oz (0.227 kg). The ranges of chemical concentrations presented are conservative,
e.g., the 12-meal-per-month levels represent the concentrations associated with 12 to 15.9 meals.
b Chronic, systemic effects
c Cancer values represent tissue concentrations at a 1 in 100,000 risk level.
Notes:
1. Consumption limits are based on an adult body weight of 70 kg and an RfD of 7x10"4 mg/kg-d.
2. None = No consumption recommended.
3. In cases where >16 meals per month are consumed, refer to Equations 3-1 and 3-2, Section 3.2.1.2, for methods
to determine safe consumption limits.
4. The detection limit for diazinon is 2x 10"3 mg/kg.
5. Instructions for modifying the variables in this table are found in Section 3.3.
6. Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD). When the
monthly limit is consumed in less than 1 month (e.g., in a few large meals), the daily dose may exceed the RfD
(see Section 2.3).
4-20
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-19. Monthly Fish Consumption Limits for Noncarcinogenic Health Endpoint -
Disulfoton
Risk Based Consumption Limit3
Fish Meals/Month
Unrestricted (>16)
16
12
8
4
3
2
1
0.5
None (<0.5)
Noncancer Health Endpointsb
Fish Tissue Concentrations
(ppm, wet weight)
0-0.012
>0.012- 0.023
>0.023- 0.031
>0.031 - 0.047
>0.047 - 0.094
>0. 094 -0.13
>0.13-0.19
>0. 19 -0.38
>0.38-0.75
>0.75
a The assumed meal size is 8 oz (0.227 kg). The ranges of chemical concentrations presented are conservative,
e.g., the 12-meal-per-month levels represent the concentrations associated with 12 to 15.9 meals.
b Chronic, systemic effects.
Notes:
1. Consumption limits are based on an adult body weight of 70 kg and an RfD of 4x10"5 mg/kg-d.
2. None = No consumption recommended.
3. In cases where >16 meals per month are consumed, refer to Equations 3-1 and 3-2, Section 3.2.1.2, for methods
to determine safe consumption limits.
4. The detection limit for disulfoton is 2 x 10"3 mg/kg.
5. Instructions for modifying the variables in this table are found in Section 3.3.
6. Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD). When the
monthly limit is consumed in less than 1 month (e.g., in a few large meals), the daily dose may exceed the RfD
(see Section 2.3).
4-21
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-20. Monthly Fish Consumption Limits for Noncarcinogenic Health Endpoint -
Ethion
Risk Based Consumption Limit3
Fish Meals/Month
Unrestricted (>16)
16
12
8
4
3
2
1
0.5
None (<0.5)
Noncancer Health Endpointsb
Fish Tissue Concentrations
(ppm, wet weight)
0-0.15
>0. 15 -0.29
>0.29-0.39
>0. 39 -0.59
>0.59-1.2
>1.2- 1.6
>1.6-2.3
>2.3-4.7
>4.7-9.4
>9.4
a The assumed meal size is 8 oz (0.227 kg). The ranges of chemical concentrations presented are conservative,
e.g., the 12-meal-per-month levels represent the concentrations associated with 12 to 15.9 meals.
b Chronic, systemic effects.
Notes:
1. Consumption limits are based on an adult body weight of 70 kg and an RfD of 5x10"4 mg/kg-d.
2. None = No consumption recommended.
3. In cases where >16 meals per month are consumed, refer to Equations 3-1 and 3-2, Section 3.2.1.2, for methods
to determine safe consumption limits.
4. The detection limit for ethion is 2 x 10"3 mg/kg.
5. Instructions for modifying the variables in this table are found in Section 3.3.
6. Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD). When the
monthly limit is consumed in less than 1 month (e.g., in a few large meals), the daily dose may exceed the RfD
(see Section 2.3).
4-22
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-21. Monthly Fish Consumption Limits for Noncarcinogenic Health Endpoint -
Terbufos
Risk Based Consumption Limit3
Fish Meals/Month
Unrestricted (>16)
16
12
8
4
3
2
1
0.5
None (<0.5)
Noncancer Health Endpointsb
Fish Tissue Concentrations
(ppm, wet weight)
0 - 0.0059
>0.0059- 0.012
>0.012- 0.016
>0.016- 0.023
>0.023 - 0.047
>0.047 - 0.063
>0.063 - 0.094
>0. 094 -0.19
>0.19-0.38
>0.38
a The assumed meal size is 8 oz (0.227 kg). The ranges of chemical concentrations presented are conservative,
e.g., the 12-meal-per-month levels represent the concentrations associated with 12 to 15.9 meals.
b Chronic, systemic effects.
Notes:
1. Consumption limits are based on an adult body weight of 70 kg and an RfD of 2x10"5 mg/kg-d.
2. None = No consumption recommended.
3. In cases where >16 meals per month are consumed, refer to Equations 3-1 and 3-2, Section 3.2.1.2, for methods
to determine safe consumption limits.
4. The detection limit for terbufos is 2 x 10"3 mg/kg.
5. Instructions for modifying the variables in this table are found in Section 3.3.
6. Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD). When the
monthly limit is consumed in less than 1 month (e.g., in a few large meals), the daily dose may exceed the RfD
(see Section 2.3).
4-23
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-22. Monthly Fish Consumption Limits for Carcinogenic and Noncarcinogenic
Health Endpoints - Oxyfluorfen
Risk Based Consumption Limit3
Fish Meals/Month
Unrestricted (>16)
16
12
8
4
3
2
1
0.5
None (<0.5)
Noncancer Health Endpointsb
Fish Tissue Concentrations
(ppm, wet weight)
0-0.88
>0.88-1.8
>1.8-2.3
>2.3-3.5
>3.5-7
>7-9.4
>9.4-14
>14-28
>28 - 56
>56
Cancer Health Endpoints0
Fish Tissue Concentrations
(ppm, wet weight)
0-0.04
>0. 04 -0.08
>0.08-0.11
>0.11 -0.16
>0.16-0.32
>0.32 - 0.43
>0.43 - 0.64
>0.64- 1.3
>1.3-2.6
>2.6
a The assumed meal size is 8 oz (0.227 kg). The ranges of chemical concentrations presented are conservative, e.g., the
12-meal-per-month levels represent the concentrations associated with 12 to 15.9 meals.
b Chronic, systemic effects.
c Cancer values represent tissue concentrations at a 1 in 100,000 risk level.
Noted:
1. Consumption limits are based on an adult body weight of 70 kg, an RfD of 3x10"3 mg/kg-d, and a cancer slope factor
(CSF) of 0.0732 (mg/kg-d)'1
2. None = No consumption recommended.
3. In cases where >16 meals per month are consumed, refer to Equations 3-1 and 3-2, Section 3.2.1.2, for methods to
determine safe consumption limits.
4. The detection limit for oxyfluorfen is 1 x 10'2 mg/kg.
5. Instructions for modifying the variables in this table are found in Section 3.3.
6. Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD). When the monthly limit
is consumed in less than 1 month (e.g., in a few large meals), the daily dose may exceed the RfD (see Section 2.3).
4-24
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-23. Monthly Fish Consumption Limits for Carcinogenic Health Endpoint - PAHs
Risk Based Consumption Limit3
Fish Meals/Month
Unrestricted (>16)
16
12
8
4
3
2
1
0.5
None (<0.5)
Noncancer Health Endpointsb
Fish Tissue Concentrations
(ppm, wet weight)
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
Cancer Health Endpoints0
Fish Tissue Concentrations
(ppm, wet weight)
0 - 0.0004
>0.0004 - 0.0008
>0.0008- 0.0011
>0.0011 -0.0016
>0.0016- 0.0032
>0.0032 - 0.0043
>0.0043 - 0.0064
>0. 0064 -0.01 3
>0.013-0.026
>0.026
a The assumed meal size is 8 oz (0.227 kg). The ranges of chemical concentrations presented are conservative, e.g., the
12-meal-per-month levels represent the concentrations associated with 12 to 15.9 meals.
b Chronic, systemic effects. An RfD is not available (NA) for this compound.
c Cancer values represent tissue concentrations at a 1 in 100,000 risk level.
Notes:
1. Consumption limits are based on an adult body weight of 70 kg and a cancer slope factor (CSF) of 7.3 (mg/kg-d)"1. No
RfD was available (June 1999).
2. None = No consumption recommended.
3. In cases where >16 meals per month are consumed, refer to Equations 3-1 and 3-2, Section 3.2.1.2, for methods to
determine safe consumption limits.
4. The detection limit for PAHs is 1 x 10"6 mg/kg.
5. Instructions for modifying the variables in this table are found in Section 3.3.
6. Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD). When the monthly limit
is consumed in less than 1 month (e.g., in a few large meals), the daily dose may exceed the RfD (see Section 2.3).
4-25
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-24. Monthly Fish Consumption Limits for Carcinogenic and Noncarcinogenic
Health Endpoints - PCBs
Risk Based Consumption Limit3
Fish Meals/Month
Unrestricted (>16)
16
12
8
4
3
2
1
0.5
None (<0.5)
Noncancer Health Endpointsb
Fish Tissue Concentrations
(ppm, wet weight)
0-0.0059
>0. 0059 -0.012
>0. 012 -0.016
>0.016- 0.023
>0.023 - 0.047
>0.047 - 0.063
>0.063 - 0.094
>0.094-0.19
>0. 19 -0.38
>0.38
Cancer Health Endpoints0
Fish Tissue Concentrations
(ppm, wet weight)
0-0.0015
>0. 0015 -0.0029
>0.0029 - 0.0039
>0.0039 - 0.0059
>0.0059- 0.012
>0.012- 0.016
>0.016- 0.023
>0.023 - 0.047
>0.047 - 0.094
>0.094
a The assumed meal size is 8 oz (0.227 kg). The ranges of chemical concentrations presented are conservative, e.g., the
12-meal-per-month levels represent the concentrations associated with 12 to 15.9 meals.
b Chronic, systemic effects
c Cancer values represent tissue concentrations at a 1 in 100,000 risk level.
* Concentration reported in parts per quadrillion (nanogram per kg or 10-9 g/kg.
Notes:
1. Consumption limits are based on an adult body weight of 70 kg, and RfD of 2x10"5, and a cancer slope factor (CSF) of 2
(mg/kg-d)-1.
2. NONE = No consumption recommended.
3. In cases where >16 meals per month are consumed, refer to Equations 3-1 and 3-2, Section 3.2.1.2, for methods to
determine safe consumption limits.
4. The detection limit for PCBs (sum of Aroclors) is 2 x 10"2 mg/kg.
5. Instructions for modifying the variables in this table are found in Section 3.3.
6. Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD). When the monthly limit
is consumed in less than 1 month (e.g., in a few large meals), the daily dose may exceed the RfD (see Section 2.3).
4-26
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4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-25. Monthly Fish Consumption Limits for Carcinogenic Health Endpoint -
Dioxins/Furans
Risk Based Consumption Limit3
Fish Meals/Month
Unrestricted (>16)
16
12
8
4
3
2
1
0.5
None (<0.5)
Noncancer Health Endpointsb
Fish Tissue Concentrations
(ppm, wet weight)
NA
NA
NA
NA
NA
NA
NA
NA
NA
NA
Cancer Health Endpoints0
Fish Tissue Concentrations
(ppt*-TEQ, wet weight)
0-0.019
>0. 01 9 -0.038
>0.038-0.05
>0.05- 0.075
>0.075-0.15
>0.15-0.2
>0.2-0.3
>0.3-0.6
>0.6-1.2
>1.2
a The assumed meal size is 8 oz (0.227 kg). The ranges of chemical concentrations presented are conservative, e.g., the
12-meal-per-month levels represent the concentrations associated with 12 to 15.9 meals.
b Chronic, systemic effects. An RfD is not available (NA) for this compound.
c Cancer values represent tissue concentrations at a 1 in 100,000 risk level.
* Concentration reported in parts per trillion (nanogram per kg or 10"9 g/kg
Notes:
1. Consumption limits are based on an adult body weight of 70 kg and a cancer slope factor (CSF) of 1.56x105 (mg/kg-d)"1.
No RfD available (June 1999).
2. None = No consumption recommended.
3. In cases where >16 meals per month are consumed, refer to Equations 3-1 and 3-2, Section 3.2.1.2, for methods to
determine safe consumption limits.
4. The detection limit for dioxins/furans is 1 x 10"6 mg/kg.
5. Instructions for modifying the variables in this table are found in Section 3.3.
6. Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD). When the monthly limit
is consumed in less than 1 month (e.g., in a few large meals), the daily dose may exceed the RfD (see Section 2.3).
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5.1 INTRODUCTION
SECTION 5
TOXICOLOGICAL PROFILE SUMMARIES FOR TARGET ANALYTES
5.1 INTRODUCTION
This section presents toxicological profile summaries for the target analytes in the
same order in which they are listed in Table 1-1. Toxicity data were collected for
the target analytes from a variety of sources. Major sources used were IRIS,
HSDB, ATSDR Toxicological Profiles, the Office of Pesticide Programs
toxicological database, and recent toxicological reviews. The EPA risk values
discussed in this section were used along with exposure data (e.g., meal size and
fish contaminant concentration) to calculate the fish consumption limits provided
in Section 4. Primary literature searches and reviews were not conducted for the
development of this section due to time and resource constraints.
EPA evaluates dose-response data for chemicals of environmental concern on an
ongoing basis. However, new toxicological data are continually being generated.
Consequently, there may be recent information that is not yet incorporated into
the EPA risk values. This may be particularly relevant for developmental toxicity,
which is the subject of much current research. The toxicological summaries
provide the reader with information that can be used to calculate alternative
health-based risk values and fish consumption limits. The methods for carrying
this out are described in Sections 2 and 3.
Risk values are also provided in the individual profiles, accompanied by a
discussion of a number of toxicity studies for each target analyte, which yield
various dose-response results. These give some indication of the variability in the
types of effects and doses at which various effects were observed.
5.1.1 Categories of Information Provided for Target Analytes
Specific types of information were sought for all target analytes to address health
and risk concerns for carcinogenic, developmental, and chronic exposure (noncar-
cinogenic) effects. These include pharmacokinetics, acute and chronic toxicity,
reproductive and developmental toxicity, mutagenicity, carcinogenicity, special
susceptibilities, interactive effects, and critical data gaps. The categories of
information provided for each target analyte are listed in Table 5-1. Although the
same types of information were sought for all analytes, the information presented
for the contaminants
5-1
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5.1 INTRODUCTION
Table 5-1. Health and Toxicological Data Reviewed for Target Analytes
Category
Specific Information
Background
Pharmacokinetics
Acute toxicity
Chronic toxicity
Reproductive and
developmental toxicity
Mutagenicity
Carcinogenicity
Special susceptibilities
Interactive effects
Critical data gaps
Summary of EPA risk values
Chemical structure/group
Use and occurrence
Target tissues
Absorption
Deposition-bioaccumulation potential/half-life/body burden
Metabolism
Excretion
Susceptible subgroups
Quantitation
Susceptible subgroups
Organ systems
Animal studies-quantitation
Human studies-quantitation
Other studies-quantitation
Database quality
Susceptible subgroups
Current risk values
Organ systems
Animal studies-quantitation
Human studies-quantitation
Other studies-quantitation
Database quality
Susceptible subgroups
Current risk values
Type
Quantitation
Source
Database quality
Organ systems
Animal studies-quantitation
Human studies-quantitation
Other studies-quantitation
Database quality
Outstanding issues
Subgroups of concern
Qualitative
Quantitative
MIXTOX results
Description
Cancer slope factor and reference dose
5-2
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5.1 INTRODUCTION
varies, depending on the types of data available. Many of the analytes listed have
been recognized as environmental contaminants for a number of years and have
a fairly comprehensive toxicological database. Others have been introduced into
the environment relatively recently; consequently, only limited information is
available on these chemicals.
When a substantial amount of information was available on a contaminant, the
information included in the discussions focused on areas relevant to the toxicities
under evaluation. For example, a significant amount of pharmacokinetic data is
available for some chemicals in the ATSDR Toxicological Profiles. In this
document, most information was briefly synopsized; however, detailed information
on human milk bioconcentration was included for developmental toxicants if
lactational exposure was of concern. In addition, when the toxicological data
indicated that a particular type of information, not reported, was required for full
exploration of relevant toxic effects, additional information was identified in the
Data Gaps Section (e.g., the interaction of DDT with pharmaceutical efficacy
arising from DDT-induced increases in levels of microsomal enzymes).
The information collected is categorized by the temporal nature of the exposure
(e.g., acute, chronic). These groupings are most applicable to the standard risk
assessment methods that were employed to calculate risk values. The temporal
groupings and methods of evaluating dose-response data are briefly discussed
in Section 2, with a description of uncertainties and assumptions associated with
dose-response evaluation.
5.1.1.1 Pharmacokinetics—
A brief summary of the pharmacokinetic data is presented for many chemicals.
The information was included if it had a bearing on the development of fish
consumption limits or would be useful to the reader in evaluating the toxicological
characteristics of a chemical. For more detailed information on pharmacokinetics,
the reader is referred to the ATSDR profiles and the primary literature.
For most chemicals there was not sufficient quantitative information, such as
absorption, uptake, distribution, metabolism, excretion, and metabolite toxicity, in
the data reviewed to recommend modifications in exposure to yield an altered
internal dose. Some chemicals contained in the IRIS database have risk values
that have incorporated pharmacokinetic considerations. If additional information
relevant to quantitative risk assessment becomes available, it will be included in
future versions of this guidance document.
5.1.1.2 Acute Toxicity—
Very little acute exposure toxicity data were located that could have a quantitative
bearing on the development of fish consumption limits. A qualitative description
of acute effects is included. The minimum estimated lethal dose to humans and
a brief discussion of the acute effects are included if the data were available.
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5.1 INTRODUCTION
5.1.1.3 Chronic Toxicity—
Under the chronic exposure heading, significant effects associated with long-term
exposure are listed. These include effects on the major organs and systems: the
liver, kidney, gastrointestinal, cardiovascular, and reproductive systems. The
chronic exposure data for each analyte include a description of an RfD listed in
IRIS or obtained from other sources and the critical study serving as the basis for
that RfD, including the species tested, duration of the study, and critical effect
noted. Information is provided on any special issues concerning the critical study
or RfD (e.g., if the study is old or has very few subjects or if the confidence in the
RfD is listed as "low").
Data are also provided on effects observed in recent dose-response studies or
effects that were not the subject of the IRIS RfD critical study. This was done to
provide a more comprehensive picture of the overall toxicological nature of the
chemicals than could be obtained from reviewing the RfD critical study alone. For
most analytes, the information is primarily a qualitative description of effects. For
chemicals that have significant new toxicological data available, details are
provided on NOAELs, LOAELs, some study characteristics, and the usual
categories of uncertainty and modifying factors that should be considered for
significant studies. These are provided to give readers the option of developing
exposure limits as they deem necessary.
5.1.1.4 Redproductive and Developmental Toxicity—
Reproductive and developmental toxicity data were obtained for each target
analyte. Section 2.3.2.3 contains general information on developmental toxicity,
including definitions and special issues related to developmental toxicity.
For many chemicals, information is provided on the tendency of the chemical to
accumulate in body tissue. Many of the target analytes bioaccumulate and/or
preferentially seek fatty tissues. When such accumulation occurs, exposure
occurring prior to pregnancy can contribute to the overall maternal body burden
and result in exposure to the developing fetus. Any body burden may result in
exposure, but lipid-seeking chemicals, such as organochlorines, are often rapidly
mobilized at the onset of pregnancy and may result in elevated contaminant
exposure to the developing fetus. As a result, it may be necessary to reduce the
exposure of females of reproductive age in order to reduce their overall body
burden. For example, if a female has been exposed to methylmercury, even if
exposure is reduced during pregnancy, the outcome of that pregnancy may be
affected depending on the timing and extent of prior exposure. This is noted for
bioaccumulative analytes in the individual toxicological profiles.
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5.1 INTRODUCTION
5.1.1.5 Mutagenicity—
Although there were many reported mutagenicity bioassays for target analytes,
little in vivo mutagenicity dose-response data were located. In vivo studies are
recommended by EPA for risk assessments of suspected mutagens. A brief
summary of the results of the mutagenicity assays for the analytes is provided.
There are numerous studies available for some of the contaminants;
consequently, it was not feasible to list all results. To provide a more concise
overview of the results of greatest concern, the nature of the positive studies is
given. The direction of the majority of results is also given (e.g., primarily positive,
negative, or mixed).
5.1.1.6 Carcinogenicity—
Cancer slope factors and descriptive data were obtained primarily from IRIS,
HEAST, and OPP. Preference was given to IRIS values; however, when IRIS
values were not available, values developed by Agency program offices (e.g.,
OPP) are provided. The program office values have not necessarily undergone
the extensive interagency review required for inclusion in the IRIS database,
although many have been reviewed by scientists within and outside of EPA.
There are often insufficient studies to evaluate the carcinogenicity of a chemical.
EPA has recognized this and formalized the lack of data as classification D: "not
classifiable as to human carcinogenicity" in EPA's cancer weight of evidence
scheme (U.S. EPA, 1986a). Many target analytes fall into this category; for
others, no data were found in the sources consulted regarding their carcin-
ogenicity. For chemicals with insufficient or no data on carcinogenicity in the
databases consulted, the text under the "Carcinogenicity" heading states that:
"insufficient information is available to determine the carcinogenic status of the
chemical." This statement is used for chemicals lacking a cancer slope factor
unless an Agency-wide review has determined that there is evidence that the
chemical is not carcinogenic (i.e., an E classification as provided in IRIS, 1999).
For a complete description of EPA's weight-of-evidence classification scheme,
see EPA's Guidelines for Carcinogenic Risk Assessment (U.S. EPA, 1986a).
EPA's proposed cancer guidelines have replaced this weight-of-evidence
classification scheme with a narrative with descriptors in three categories:
"known/likely," "cannot be determined," or "not likely" (U.S. EPA, 1996b).
5.1.1.7 Special Susceptibilities—
Toxicity data often indicate that some groups of individuals may be at greater risk
from exposure to chemicals or chemical groups. For example, a chemical that
causes a specific type of organ toxicity will usually pose a greater risk to
individuals who have diseases of that organ system (e.g., immunotoxicity poses
a greater risk to those with immunosuppression or with immature immune
systems). Persons with some genetic diseases (e.g., enzyme disorders),
nutritional deficiencies, and metabolic disorders may also be at greater risk due
5-5
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5.1 INTRODUCTION
to exposure to some chemicals. Qualitative data on special susceptibilities are
provided for many of the target analytes. However, there are no quantitative data
on subgroup susceptibilities for most chemicals that would enable the risk
assessor to modify risk values.
The RfDs are designed to take into account the most susceptible individuals, and
RfDs often incorporate an uncertainty factor to account for variability within the
human species. Susceptible subgroups are those that exhibit a different or more
enhanced response than most persons exposed to the same level of the chemical
in the environment. Reasons include genetic makeup, developmental stage, age,
health and nutritional status (including dietary habits that may increase
susceptibility, such as inconsistent diets or nutritional deficiencies), and substance
exposure history (including smoking).
5.1.1.8 Interactive Effects-
Data on interactive effects were located for many, but not all, of the target
analytes. Most data on interactive effects were obtained from ATSDR Toxico-
logical Profiles. Often the data indicate that certain classes of chemicals may be
of concern. For example, mostorganochlorines induce the mixed function oxidase
system. These chemicals may lead to unanticipated and exaggerated or dimin-
ished effects arising from simultaneous exposure to other chemicals that rely on
the same metabolic system. In some cases this leads to potentiation (increased
toxicity) and in others it hastens the process of detoxification.
The MIXTOX database, developed by EPA, was also used to obtain information
on interactive effects (MIXTOX, 1992). The database provides a very brief sum-
mary of results of studies on combinations of chemicals. Most interactions are
reported as "potentiation," "inhibition" or "antagonism" (decreased toxicity), "no
apparent influence," or "additive." The interactions that differ from additive or no
apparent influence are reported because it is assumed, in the absence of contrary
information, that the toxicity of mixtures of chemicals will be additive for the same
target tissue (see Section 2.3). The interactive terminology used in MIXTOX is
used in this document.
5.1.1.9 Critical Data Gaps-
Data gaps noted in IRIS files, the OPP toxicological database, RfD summaries,
and the ATSDR Toxicological Profiles are listed. In addition, data gaps that have
been identified from a review of the studies are listed, along with the reasons that
additional data are considered necessary.
5.1.1.10 Summary of EPA Levels of Concern—
The EPA risk values (RfDs and cancer slope factors) discussed in each section
and used in the development of fish consumption limits are summarized in
Table 3-1.
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5.1 INTRODUCTION
5.1.1.11 Major Sources—
At the end of each target analyte file is a list of the major sources of information
consulted. Major sources are those that have been cited more than once. Within
the text of each target analyte file, all information is provided with citations.
The IRIS files were consulted in early 1999 for cancer slope factors, chronic
exposure Rf Ds, and additional study data. ATSDR Toxicological Profiles were also
consulted when available. The profiles have extensive toxicity, pharmacokinetic,
and epidemiological data reviews.
5.1.1.12 Statement Regarding Uncertainty-
There are always significant uncertainties associated with estimating health risks
and safe exposure levels for human populations. Although these are discussed
in Section 2, their importance warrants their mention in this section also. The risk
values provided for each chemical in this section are based on human or animal
studies that evaluated either a small subset of the human population or an entirely
different species. In either case, we can only estimate the relevance of the study
results to humans. Although a quantitative methodology is used to extrapolate
from various types of studies to the general human population, there is consider-
able uncertainty in the estimated relationship between study populations and the
human population.
The use of uncertainty factors and upper bound cancer risk estimates provides a
margin of safety to account for some aspects of uncertainty in the extrapolation.
However, our knowledge of response variability in the human population is very
limited. The variations in response, which are engendered by age, sex, genetic
heterogeneity, and preexisting disease states, may be considerable. Con-
sequently, although current approaches to assessing risk involve estimating the
upper bound values for deriving exposure or risk and are intended to be protective
rather than predictive, the reader is urged to carefully review the information
provided in this section on data gaps and uncertainties.
It is important to describe the uncertainties and assumptions when recommending
fish consumption limits. With respect to toxicity, these include both uncertainties
associated with specific chemicals and uncertainties and assumptions associated
with the dose-response evaluation process (described in Section 2). In some
cases, a variety of dose-response data will enable the reader to provide a
quantitative estimation of the range of potential risk values that could be used to
calculate exposure and fish consumption limits. A description of data gaps may
also be useful to the risk manager in determining the best course of action. For
chemicals having limited data, only a qualitative description may be possible.
5-7
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5.1 INTRODUCTION
5.1.2 Abbreviations Used and Scientific Notation
The glossary contains a description of additional terms and abbreviations used in
this section.
Scientific notation is used where the values are less than 0.001 unless it would
introduce confusion to the text (e.g., when presenting a range, the same format
is used for both values in the range). In the summaries of risk values, all
noncancer risk values are presented in scientific notation to facilitate comparison
across health endpoints.
5-8
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5.2 METALS
5.2 METALS
5.2.1 Arsenic
5.2.1.1 Background-
Arsenic is a naturally occurring element in the earth's crust that is usually found
combined with other elements. Arsenic combined with elements such as oxygen,
chlorine, and sulfur is referred to as inorganic arsenic; arsenic combined with
carbon and hydrogen is referred to as organic arsenic. In this toxicological profile,
arsenic refers to inorganic arsenic and its associated compounds. Organic
arsenic compounds, such as arsenobetaine (an organic arsenic compound found
in the edible parts of fish and shellfish) are not discussed, since these compounds
are considered to be relatively nontoxic and not a threat to human health (ATSDR,
1999c).
5.2.1.2 Pharmacokinetics—
Pharmacokinetic studies show that water-soluble arsenic compounds are well-
absorbed across the gastrointestinal tract. They appear to be transported
throughout the body. Analysis of tissues taken at autopsy from people who were
exposed to arsenic found arsenic present in all tissues of the body. The arsenic
levels in hair and nails were the highest, with somewhat lower levels in internal
organs (ATSDR, 1999c).
The metabolism of arsenic consists mainly of a reduction reaction, which converts
pentavalent arsenic to trivalent arsenic, and methylation reactions, which convert
arsenite to monomethylarsonic acid and dimethylarsenic acid. The primary
excretion route for arsenic and metabolites is in the urine, with human studies
showing that 45 to 85 percent is excreted in the urine within 1 to 3 days of
ingestion. Very little is excreted in the feces (ATSDR, 1999c).
5.2.1.3 Acute Toxicity—
Arsenic is a recognized human poison. Single large doses, approximately 600
ug/kg-d or higher, taken orally have resulted in death. Acute oral exposure to
lower levels of arsenic has resulted in effects on the gastrointestinal system
(nausea, vomiting, diarrhea); central nervous system (headaches, weakness,
lethargy, delirium); cardiovascular system (sinus tachycardia, hypotension,
shock); and the liver, kidney, and blood (anemia, leukopenia). The limited
available data have shown arsenic to have low to moderate acute toxicity to
animals. Lethal oral doses to animals are higher than those in humans based on
data showing that the oral LD50 values for arsenic range between 15 and 112
mg/kg (ATSDR, 1999c).
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5.2 METALS
5.2.1.4 Chronic Toxicity—
The primary effects noted in humans from chronic exposure to arsenic are effects
on the skin. Oral exposure has resulted in a pattern of skin changes that include
the formation of warts or corns on the palms and soles along with areas of
darkened skin on the face, neck, and back. Blackfoot disease, a disease
characterized by a progressive loss of circulation in the hands and feet, leading
ultimately to necrosis and gangrene, is associated with arsenic exposure (ATSDR,
1999c). Other effects noted from chronic oral exposure include peripheral
neuropathy, cardiovascular disorders, gastrointestinal disorders, hematological
disorders, and liver and kidney disorders.
IRIS provides an RfD for inorganic arsenic of 3.0 x 10~4 mg/kg-d, based on a
NOAEL (adjusted to include arsenic exposure from food) of 0.0008 mg/kg-d and
an uncertainty factor of 3. This was based on two studies that showed that the
prevalence of hyperpigmentation and skin lesions increased with both age and
dose for individuals exposed to high levels of arsenic in drinking water. There
were also some cardiovascular effects noted. Other human studies support these
findings, with several studies noting an increase in skin lesions from chronic
exposure to arsenic through the drinking water. An uncertainty factor of 3 was
used to account for both the lack of data to preclude reproductive toxicity as a
critical effect and for uncertainty as to whether the NOAEL of the critical studies
accounts for all sensitive individuals (IRIS, 1999).
EPA has medium confidence in the studies on which the RfD was based and in
the RfD. The key studies were extensive epidemiologic reports that examined the
effects of arsenic in a large number of people. However, doses were not well-
characterized, othercontaminants were present, and potential exposure from food
or other sources was not examined. The supporting studies suffer from other
limitations, primarily the small populations studied. However, the general
database on arsenic does support the findings in the key studies; this was the
basis for EPA's "medium confidence" ranking of the RfD (IRIS, 1999).
5.2.1.5 Reproductive and Developmental Toxicity—
Limited information is available on the developmental effects of arsenic in
humans. No overall association between arsenic in drinking water and congenital
heart defects was detected in an epidemiological study, although an association
with one specific lesion (coarctation of the aorta) was noted. In another study, a
marginal association (not statistically significant) was found between detectable
levels of arsenic in drinking water and spontaneous abortions. The odds ratio for
the group with the highest arsenic concentration was statistically significant.
However, a similar association was found for a number of compounds, which
indicates that the association could be random or due to other risk factors
(ATSDR, 1999c). A study of babies born to women exposed to arsenic dusts in
a copper smelter in Sweden showed a higher-than-expected incidence of
congenital malformations.
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5.2 METALS
Minimal or no effects on fetal development have been observed in chronic oral
exposure studies of pregnant rats or mice to low levels of arsenic in drinking
water. Malformations were produced in 15-d hamster fetuses via intravenous
injections of arsenic into pregnant dams on day 8 of gestation, while another study
reported that very high single oral doses of arsenic were necessary to cause
prenatal fetal toxicity (IRIS, 1999).
5.2.1.6 Mutagenicity—
Arsenic has not been reported to directly react with DMA in many studies. Studies
have shown that arsenic chromosomal aberrations and sister chromatid exchange
in human lymphocytes reported positive results, while others were negative. One
study in mouse bone marrow cells reported an increase in micronuclei, while
another did not report an increase in chromosomal breaks and exchanges
(ATSDR, 1999c). In vitro studies have also reported both positive and negative
results. Arsenic was negative in the bacterial colorimetric assay: SAS Chromotest
(HSDB, 1999), and positive for reverse mutations in bacteria, morphological
transformations, sister chromatid exchange, and chromosomal aberrations in
Syrian hamster embryo cells. Arsenic was also positive for chromosomal
aberrations in human leukocytes and lymphocytes, sister chromatid exchange,
enhancement or inhibition of DMA synthesis, and hyperdiploidy and chromosomal
breakage in human lymphocytes (ATSDR, 1999c).
5.2.1.7 Carcinogenicity—
EPA has classified inorganic arsenic in Group A—Known Human Carcinogen.
This is based on the increased incidence in humans of lung cancer through
inhalation exposure and the increased risk of nonmelanoma skin, bladder, liver,
kidney, and lung cancer through drinking water exposure (IRIS, 1999).
Animal studies have not associated arsenic exposure, via ingestion, with cancer.
All cancer studies in rodents with arsenic have reported negative results.
However, the meaning of this nonpositive data is uncertain because the
mechanism of action in causing human cancer is not known, and rodents may not
be good models for arsenic-induced carcinogenicity (IRIS, 1999).
To estimate the risks posed by ingestion of arsenic, EPA uses data from Taiwan
concerning skin cancer incidence, age, and level of exposure via drinking water.
In 37 villages that had obtained drinking water for 45 years from artesian wells
with various elevated levels of arsenic, more than 40,000 individuals were
examined for hyperpigmentation, keratosis, skin cancer, and blackfoot disease.
The local well waters were analyzed for arsenic, and the age-specific cancer
prevalence rates were found to be correlated with both local arsenic
concentrations and age (duration of exposure). The oral cancer potency is 1.5
per mg/kg-d (IRIS, 1999).
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5.2 METALS
EPA's current regulation for arsenic in drinking water (50 |ig/L) has recently been
called into question. The conclusions of a recent National Research
Council/National Academy of Sciences report on arsenic in drinking water suggest
that the current drinking water regulation needs to be lowered based on risks of
skin, lung, and bladder cancer (NRC, 1999).
5.2.1.8 Special Susceptibilities-
No studies regarding unusual susceptibility of any human subpopulation to arsenic
are available. However, it is possible that some members of the population might
be especially susceptible because of lower than normal methylating capacity.
This could result from a dietary deficiency of methyl donors such as choline or
methionine or a deficiency of the vitamin coenzymes (folacin, Vitamin B12)
involved in transmethylation reactions (ATSDR, 1999c; Rogers, 1995).
5.2.1.9 Interactive Effects-
Arsenic tends to reduce the effects of selenium, and selenium can decrease the
effects of arsenic. No clear evidence exists for significant interactions between
arsenic and other metals; the existing data do not suggest that arsenic toxicity is
likely to be significantly influenced by concomitant exposure to other metals.
Some evidence suggests that a positive interaction between arsenic and
benzo(a)pyrene can occur for lung adenocarcinomas in animals. Other studies
suggest that chemicals that interfere with the methylation process could increase
the toxicity of arsenic (ATSDR, 1999c)
5.2.1.10 Critical Data Gaps-
There is a substantial database on the toxicity of arsenic, both in humans and in
animals. However, there are some areas where studies are lacking. Further
epidemiological studies on the health effects of arsenic at low doses would be
valuable. Additional studies on developmental and reproductive effects of arsenic
would also be useful (ATSDR, 1999c).
5.2.1.11 Summary of EPA Health Benchmarks-
Chronic Toxicity 3.0 x 10~4 mg/kg-d
Carcinogenicity 1.5 per mg/kg-d.
5.2.1.12 Major Sources—
ATSDR (1999c), HSDB (1999), IRIS (1999), Rogers (1995).
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5.2 METALS
5.2.2 Cadmium
5.2.2.1 Background-
Cadmium is a heavy metal that occurs naturally in the earth's crust. It can be
released into the environment through a wide variety of industrial and agricultural
activities. It accumulates in human and other biological tissue and has been
evaluated in both epidemiological and toxicological studies. ATSDR has
determined that exposure conditions of most concern are long-term exposures to
elevated levels in the diet (ATSDR, 1997).
The FDA has estimated that cadmium exposure among smokers is approximately
10 ug/d (0.01 mg/d). Passive exposure of nonsmokers may also be a source of
exposure (U.S. FDA, 1993). This should be considered in evaluating the total
exposure and risks associated with cadmium.
5.2.2.2 Pharmacokinetics—
Cadmium is not readily absorbed when exposure occurs via ingestion. Most
ingested cadmium passes through the gastrointestinal (Gl) tract without being
absorbed. Studies in humans indicate that approximately 25 percent of cadmium
consumed with food was retained in healthy adults after 3 to 5 days; this value fell
to 6 percent after 20 days. Absorption may be much higher in iron-deficient
individuals. Evaluations of the impact of cadmium complexation indicate that
cadmium absorption from food is not dependent upon chemical complexation.
However, some populations with high dietary cadmium intakes have elevated
blood cadmium levels, which may be due to the particular forms of cadmium in
their food (ATSDR, 1997).
Cadmium absorption studies in animals indicate that the proportion of an oral
dose that is absorbed is lower in animals than in humans. Absorption is elevated
during pregnancy, with whole-body retention in mice of 0.2 percent in those that
had undergone pregnancy and lactation and 0.08 percent in those that had not.
In rats, absorption decreased dramatically over the early lifetime, ranging from 12
percent at 2 hours to 0.5 percent at 6 weeks after birth. The placenta may act as
a partial barrier to fetal exposure, with cord blood concentrations being
approximately half those of maternal blood. The human data on placental
concentrations are conflicting. Cadmium levels in human milk are approximately
5 to 10 percent of those found in blood (ATSDR, 1997).
Much of the cadmium absorbed into the blood is sequestered by metallothionein,
and plasma cadmium is found primarily bound to this protein. This binding
appears to protect the kidney from the otherwise toxic effects of cadmium. It has
been suggested that kidney damage by cadmium occurs primarily due to unbound
cadmium (ATSDR, 1997). Once cadmium is absorbed, it is eliminated slowly; the
biological half-life has been estimated at 10 to 30 years (U.S. FDA, 1993).
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5.2 METALS
Body stores of iron, zinc, and calcium may affect absorption and retention,
although the retention may not be in readily available tissues (e.g., intestinal wall
versus blood). The greatest concentrations of cadmium are typically found in the
liver and kidney. Cadmium is not directly metabolized, although the cadmium ion
binds to anionic groups in proteins, especially albumin and metallothionein
(ATSDR, 1997).
5.2.2.3 Acute Toxicity—
Effects of acute oral exposure to cadmium include Gl irritation, nausea, vomiting,
abdominal pain, cramps, salivation, and diarrhea. In two human cases, lethal
doses caused massive fluid loss, edema, and widespread organ destruction. The
ingested doses that caused death were 25 mg cadmium/kg and 1,840 mg
cadmium/kg (ATSDR, 1997).
5.2.2.4 Chronic Toxicity—
Kidney toxicity is a significant concern with cadmium exposure. Increased death
rates from renal disease have been observed in exposed human populations in
Belgium, England, and Japan (ATSDR, 1997). There are also extensive animal
data indicating that the kidney is a target organ. IRIS contains an RfD of 0.001
mg/kg-d in food based upon a NOAEL of 0.01 mg/kg-d in multiple human studies.
The critical effect was significant proteinuria (an indicator of kidney toxicity). To
calculate the RfD, it was assumed that 2.5 percent of cadmium in food was
absorbed and approximately 5 percent in water was absorbed. Using an
uncertainty factor of 10 to account for intrahuman variability in cadmium
sensitivity, the RfD for cadmium in food was calculated to be 0.001 mg/kg-d. The
RfD was calculated using a toxicokinetic model to determine the highest level of
cadmium in the human renal cortex not associated with significant proteinuria and
therefore was not based on a single study. EPA's confidence in the database and
the RfD is high (IRIS, 1999).
The FDA has calculated a tolerable daily intake of 55 ug/person-d, which is
approximately equal to 0.78 ug/kg-d (7.8 x 10~4 mg/kg-d) in a 70-kg person and
5.5 ug/kg-d (0.005 mg/kg-d) in a 10-kg child (their example uses 2+ years of age).
The FDA value is based upon a pharmacokinetic approach that utilized the critical
body burden associated with kidney toxicity. See U.S. FDA (1993) for more
details.
Cadmium causes many other types of toxic effects in addition to nephrotoxicity.
In humans, some studies have suggested an association between neurotoxicity
and cadmium exposure at levels below those that cause kidney toxicity (no
additional details available). Cadmium exposure reduces the Gl uptake of iron,
which may cause anemia if iron intakes are low. Bone disorders including
osteomalacia, osteoporosis, and spontaneous bone fracture have been observed
in some chronically exposed individuals. Increased calcium excretion associated
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with cadmium-induced renal damage may lead to increased risk of osteoporosis,
especially in postmenopausal women, many of whom are already at risk of
osteoporosis. Cardiovascular toxicity and elevated blood pressure have been
suggested in some human studies; however, the results are conflicting (ATSDR,
1997).
Animal studies indicate that cadmium ingestion causes a wide variety of
alterations in the function of the immune system. Some aspects of the system
were enhanced and others were impaired (e.g., susceptibility to virally induced
leukemia). In short-term studies, serious effects occurred at levels as low as 1.9
mg/kg-d and less serious effects (induction of antinuclear antibodies) at 0.75
mg/kg-d in a 10-wk study in mice (ATSDR, 1997). No longer-term studies were
located.
5.2.2.5 Reproductive and Developmental Toxicity—
Reproductive and developmental toxicity has been associated with oral cadmium
exposure both in short- and long-term studies. In 10-d prenatal dosing studies in
rats at 18.4 mg/kg, malformations, including split palate and dysplasia of the facial
bones and rear limb, were observed with a NOAEL of 6.1 mg/kg-d. A similar
study in rats found delayed ossification at 2 mg/kg-d. Other studies have found
gross abnormalities and reduced fetal weight at doses ranging from 1.5 to 19.7
mg/kg-d (ATSDR, 1997). Oral cadmium exposure of young mice depresses their
humoral immune responses; the study did not find the same effect in adult mice
(ATSDR, 1997).
More sensitive measures of effects for cadmium have identified effects at much
lower doses. ATSDR has determined that:
... the most sensitive indicator of development toxicity of cadmium in
animals appears to be neurobehavioral development. Offspring of female
rats orally exposed to cadmium at a dose of 0.04 mg/kg-day prior to and
during gestation had reduced exploratory locomotor activity and rotorod
performance at age 2 months. . . (ATSDR, 1997).
Reduced locomotor activity and impaired balance were noted at a LOAEL of 0.04
mg/kg-d with 11 weeks of exposure occurring prior to and during gestation. The
effects were also observed at 0.7 mg/kg-d with exposure occurring only during
gestation. Neurobehavioral effects were observed in other developmental studies
and in chronic studies of effects in adult animals. Two longer-term studies
yielding similar neurobehavioral results were conducted with maternal exposures
of 7.0 and 14.0 mg/kg-d (see numerous citations in Baranskietal., 1983; ATSDR
1997).
Studies of developmental toxicity in human populations have been conducted on
women exposed via inhalation in the workplace. Decreased birth weight has been
reported in two studies, one with statistically significant results and the other
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lacking statistical significance. Inhalation studies in animals have found structural
and neurobehavioral abnormalities similar to those found in the oral dosing
studies (ATSDR, 1997).
Based on the mutagenicity data results (discussed below), heritable defects may
result from exposure to cadmium. However, mutagenicity assays do not provide
dose-response data suitable for use for the calculation of a risk value. Calcium
deficiency has been shown to increase the fetotoxicity of cadmium, and lindane
exposure increased developmental toxicity in animal studies (ATSDR, 1997).
5.2.2.6 Mutagenicity—
Results of bacteria and yeast assays have been mixed. Results were conflicting
in chromosomal aberration studies on human lymphocytes treated both in vitro
and obtained from exposed workers. Mouse and hamster germ cell studies
indicate that cadmium may interfere with spindle formation resulting in aneuploidy.
Positive results have also been obtained in Chinese hamster ovary and mouse
lymphoma cell assays (IRIS, 1999).
5.2.2.7 Carcinogenicity—
Epidemiological studies have been conducted on population groups in high
cadmium exposure areas via food and water, and organ-specific cancer rates
have been examined (kidney, prostate, and urinary tract). Most studies yielded
negative results. A study in Canada found that elevated rates of prostate cancer
paralleled the elevated cadmium exposure of the populations studied. In animals,
oral studies conducted at relatively low exposure levels (up to 4.4 mg/kg-d) have
yielded negative results. One study in rats showed an increase in prostatic
proliferative lesions, leukemia, and testicular tumors in rats fed cadmium in a zinc-
controlled diet. Rats fed zinc-deficient diets had decreased overall incidence for
tumors of the prostate, testes, and hematopoietic system thus indicating that zinc
deficiency in the diet may inhibit the carcinogenic effects of cadmium ingestion.
EPA has determined that data are insufficient to determine the carcinogenic
status of cadmium by the oral route.
An increased risk for respiratory tract cancers has been observed in several
epidemiological studies of workers exposed to cadmium-containing fumes and
dusts. For this reason, cadmium is classified as a probable human carcinogen
(B1) by EPA based on inhalation studies in humans. The airborne cancer
potency is 1.8 x 10"3 per ug/m3 (IRIS, 1999).
5.2.2.8 Special Susceptibilities-
Populations with genetically determined lower ability to induce metallothionein are
less able to sequester cadmium. Populations with depleted stores of dietary
components such as calcium and iron due to multiple pregnancies and/or dietary
deficiencies may have increased cadmium absorption from the Gl tract.
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Increased calcium excretion associated with cadmium-induced renal damage may
lead to increased risk of osteoporosis, especially in postmenopausal women. The
relationship between cadmium toxicity and iron levels is not well established;
however, in some studies it appears that iron-deficient individuals may be at
greater risk. Individuals with kidney disease, diabetes, and age-related decreased
kidney function may be at greater risk of cadmium-induced kidney toxicity
(ATSDR, 1997).
Immunological effects may be of concern for children because it appears, based
upon animal studies, that young individuals may be at greater risk than adults. In
addition, the immune system is not fully developed in humans until approximately
12 years of age. Immunological effects have also been observed in multiple
animal studies of adults. These pose special risks for individuals with
compromised immune systems (e.g., those with AIDS).
A variety of types of developmental effects have been associated with cadmium
exposure (see discussion above). These all pose special risks for infants and
children, as well as women of reproductive age.
5.2.2.9 Interactive Effects-
Dietary deficiencies of calcium, protein, zinc, copper, iron, and vitamin D may
cause increased susceptibility to adverse skeletal effects from cadmium exposure.
Lead increased neurotoxicity and selenium decreased the clastogenic effect of
cadmium on bone marrow. Exposure to chemicals that induce metallothionein
(e.g., metals) reduced toxicity with parenteral cadmium exposure (ATSDR, 1997).
MIXTOX reports a number of interactive studies on cadmium and selenium
compounds. The studies have yielded mixed results with reports of inhibition,
potentiation, additive effects, and no effects (MIXTOX, 1992).
5.2.2.10 Critical Data Gaps—
A joint team of scientists from ATSDR, National Toxicology Program (NTP), and
EPA have identified the following data gaps: immunotoxicity, neurotoxicity, and
developmental toxicity in human populations, quantitative data on acute and
intermediate toxicity in humans, and chronic exposure studies in humans using
sensitive indicators of kidney toxicity, animal and human studies of carcinogenic
effects, human genotoxicity, animal reproductive, immunotoxicity, and pharma-
cokinetic studies (ATSDR, 1997).
5.2.2.11 Summary of EPA Health Benchmarks-
Chronic Toxicity 1 x I0~3mg/kg-d
Carcinogenicity Group B1 (probable human carcinogen).
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5.2.2.12 Major Sources—
ATSDR (1997), HSDB (1993), IRIS (1999), U.S. FDA (1993).
5.2.3 Mercury
5.2.3.1 Background-
Mercury is widely distributed in the environment due to both natural and
anthropogenic processes. It is released generally as elemental mercury (Hg°) or
divalent mercury (Hg2+). It can be converted between these forms and may form
mercury compounds by chemical processes in air, water, and soil. Biological
processes in other media, primarily soil and sediment, can convert inorganic
mercury into organic, mostly methylmercury.
In fish tissue, the majority of mercury is methylmercury. Generally, the amount
of mercury in fish tissue increases with the age and the size of the fish. The
accumulation of mercury in fish varies among species; for the most part, the fish-
eating species offish accumulate higher concentrations of mercury than do non-
piscivorous fish. Mercury is found in highest concentrations in organs and
muscle.
Data on mercury toxicity have been reviewed for inclusion in IRIS. Currently there
are both RfDs and cancer assessments in IRIS for elemental mercury, inorganic
mercury (mercuric chloride), and methylmercury (interim RfD). EPA, in response
to a mandate of the Clean Air Act Amendments of 1990, has prepared a
multivolume Mercury Study Report to Congress. This has been peer reviewed
extensively including a recent review by the Science Advisory Board (SAB). (U.S.
EPA, 1997d). Methylmercury has also been the subject of evaluation by
numerous states. Detailed analyses have been conducted in some specific areas,
including evaluation of data regarding blood and hair mercury levels, toxic effects,
and biological half-life values to estimate safe consumption levels of contaminated
fish (Shubat, 1991, 1993; Stern, 1993).
As discussed in previous sections, a total exposure assessment is beyond the
scope of this document. Readers may wish to consult other sources to obtain
information on background levels of methylmercury in the environment. Additional
information on dietary sources of mercury is available in the FDA Adult Total Diet
Study, conducted from October 1977 through September 1978, which contains
information on total mercury content (not restricted to methylmercury) in a number
of foods (Podrebarac, 1984). Readers are also referred to Volume III, An
Assessment of Exposure from Anthropogenic Mercury Emissions in the United
States of the Mercury Study Report to Congress (U.S. EPA, 1997d).
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5.2.3.2 Pharmacokinetics—
Methylmercury is rapidly and nearly completely absorbed from the gastrointestinal
tract; 90 to 100 percent absorption is estimated (WHO, 1990).
Methylmercury is somewhat lipophilic, allowing it to pass through lipid membranes
of cells and facilitating its distribution to all tissues, and it binds readily to proteins.
Methylmercury in fish binds to amino acids in fish muscle tissue.
The highest methylmercury levels in humans are generally found in the kidneys.
Methylmercury in the body is considered to be relatively stable and is only slowly
transformed to form other forms of mercury. Methylmercury readily crosses the
placental and blood/ brain barriers. Estimates for its half-life in the human body
range from 44 to 80 days (U.S. EPA, 1997d). Excretion of methylmercury is via
the feces, urine, and breast milk. Methylmercury is also distributed to human hair
and to the fur and feathers of wildlife; measurement of mercury in these materials
has served as a useful biomonitor of contamination levels.
5.2.3.3 Acute Toxicity—
Acute high-level exposures to methylmercury may result in impaired central
nervous system function, kidney damage and failure, gastrointestinal damage,
cardiovascular collapse, shock, and death. The estimated lethal dose is 10 to 60
mg/kg (ATSDR, 1999).
5.2.3.4 Chronic Toxicity—
Although both elemental and methylmercury produce a variety of health effects
at relatively high exposures, neurotoxicity is the effect of greatest concern. This
is true whether exposure occurs to the developing embryo or fetus during
pregnancy or to adults and children.
Human exposure to methylmercury has generally been through consumption of
contaminated food. Two major episodes of methylmercury poisoning through fish
consumption have occurred. The first occurred in the early 1950s among people,
fish-consuming domestic animals such as cats, and wildlife living near Minamata
City on the shores of Minamata Bay, Kyushu, Japan. The source of the
methylmercury contamination was effluent from a chemical factory that used
mercury as a catalyst and discharged wastes into the bay where it accumulated
in the tissues of fish and shellfish that were dietary staples of this population.
Average fish consumption was reported to be in excess of 300 g/d (reviewed by
Harada et al., 1995); 20 times greater than is typical for recreational fishers in the
United States. By comparison, about 3 to 5 percent of U.S. consumers routinely
eat 100 grams of fish per day. Among women of childbearing age, 3 percent
routinely eat 100 grams offish per day.
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Symptoms of Minamata disease in children and adults included: impairment of
peripheral vision, disturbances in sensations ("pins and needles" feelings,
numbness) usually in the hands and feet and sometimes around the mouth,
incoordination of movements as in writing, impairment of speech, hearing, and
walking, and mental disturbances. It sometimes took several years before
individuals were aware that they were developing the signs and symptoms of
methylmercury poisoning. Over the years, it became clear that nervous system
damage could occur to the fetus if the mother ate fish contaminated with
methylmercury during pregnancy.
In 1965, another methylmercury poisoning incident occurred in the area of Niigata,
Japan. The signs and symptoms of disease in Niigata were similar to those of
methylmercury poisoning in Minamata.
Methylmercury poisoning also occurred in Iraq following consumption of seed
grain that had been treated with a fungicide containing methylmercury. The first
outbreak occurred prior to 1960; the second occurred in the early 1970s.
Imported mercury-treated seed grains that arrived after the planting season were
ground into flour and baked into bread. Unlike the long-term exposures in Japan,
the epidemic of methylmercury poisoning in Iraq was short in duration lasting
approximately 6 months.
The signs and symptoms of disease in Iraq were predominantly in the nervous
system: difficulty with peripheral vision or blindness, sensory disturbances,
incoordination, impairment of walking, and slurred speech. Both children and
adults were affected. Infants born to mothers who had consumed methyl mercury-
contaminated grain (particularly during the second trimester of pregnancy)
showed nervous system damage even though the mother was only slightly
affected.
Recent studies have examined populations that are exposed to lower levels of
methylmercury as a consequence of routine consumption of fish and marine
mammals, including studies of populations around the Great Lakes and in New
Zealand (Kjellstrom et al., 1986a, 1986b), the Amazon basin (e.g., Lebel et al.,
1996; Marsh etal., 1995b), the Seychelles Islands (Marsh etal., 1995a), and the
Faroe Islands (Dahl etal., 1996). The last two studies are of large populations of
children presumably exposed to methylmercury in utero. Very sensitive measures
of developmental neurotoxicity in these populations are still being analyzed and
published. A 1998 workshop discussed these studies and concluded that they
have provided valuable new information on the potential health effects of
methylmercury. Significant uncertainties remain, however, because of issues
related to exposure, neurobehavioral end points, confounders and statistics, and
study design.
The EPA interim RfD for methylmercury is based on data on neurologic changes
in 81 Iraqi children who had been exposed in utero; that is, their mothers had
eaten methylmercury-contaminated bread during pregnancy. The data were
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collected by interviewing the mothers of the children and by clinical examination
by pediatric neurologists conducted approximately 30 months after the poisoning
episode. The incidence of several endpoints (including late walking, late talking,
seizures, or delayed mental development and scores on clinical tests of nervous
system function) were mathematically modeled to determine a mercury level in
hair (measured in all the mothers in the study) that was associated with no
adverse effects. Delays in motor and language development were defined by the
following criteria:
• Inability to walk two steps without support by 2 years of age
• Inability to respond to simple verbal communication by age 2 years among
children with good hearing
• Scores on physical examination by a neurologist who assessed cranial nerve
signs, speech, involuntary movements, limb tone, strength, deep tendon
reflexes, plantar responses, coordination, dexterity, primitive reflexes,
sensation, posture, and ability to sit, stand, walk, and run
• Assessment of mental development or the presence of seizures based on
interviews with the child's mother.
In calculating the mercury level in hair that was associated with no adverse effects
in children exposed in utero, EPA used a benchmark dose (in this instance the
lower bound for 10 percent risk of neurological changes) based on modeling of all
effects in children. This lower bound was 11 ppm methylmercury in maternal hair.
A dose-conversion equation was used to estimate a daily intake of 1.1 ug
methylmercury/kg body weight-day that, when ingested by a 60-kg individual, will
maintain a concentration of approximately 44 ug/L of blood or a hair concentration
of 11 ug mercury/g hair (11 ppm).
A composite uncertainty factor of 10 was used to account for the following:
variability in the human population (particularly the variation in biological half-life
and variability in the hair-to-blood ratio for mercury), lack of data on long-term
sequelae of exposure, and the lack of a two-generation reproductive study. The
resulting interim RfD for methylmercury is 1 x 10~4 mg/kg-d or 0.1 ug/kg-d (IRIS,
1999).
The range of uncertainty in the interim methylmercury RfD and the factors
contributing to this range were evaluated in qualitative and quantitative uncertainty
analyses. The uncertainty analyses indicated that paresthesia (numbness or
tingling) in the hands and feet and occasionally around the mouth in adults is not
the most reliable endpoint for dose-response assessment because it is subject to
the patient's recognition of the effect. Paresthesia in adults is not the basis for
EPA's interim methylmercury RfD.
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There are, however, uncertainties associated with the interim RfD based on
developmental effects from methylmercury in children exposed in utero. There
are difficulties with reliability in recording and classifying events such as late
walking in children because the data were collected approximately 30 months
after the child's birth. In addition, the data were collected on a population that did
not necessarily follow Western cultural practices or use Western calendars in the
recording of events such as first steps or first words. It should be noted, however,
that the endpoints used represented substantial developmental delays; for
example, a child's inability to walk two steps without support at 2 years of age,
inability to talk based on use of two or three meaningful words by 2 years, or
presence of generalized convulsive seizures. There is both variability and
uncertainty in the pharmacologic parameters that were used in estimating the
ingested mercury dose. There is also a degree of uncertainty introduced by the
size of the study population (81 mother-child pairs).
The interim RfD is supported by additional studies in children exposed in utero.
These include investigations among Cree Indians in Canada and NewZealanders
who consume large amounts of fish. In these studies, the hair concentration of
mercury was used to monitor mercury exposure over time. Conclusions by the
investigators in their official reports cite developmental delays among the children
born of mothers whose hair mercury concentrations during pregnancy were 6 to
18 ppm, consistent with the benchmark dose of 11 ppm. The published data on
the pilot study portion of the ongoing work in the Seychelles Islands (data on
children of about 5 years of age) are also consistent with EPA's benchmark dose.
A 1997 review by the Science Advisory Board determined that the RfD is
scientifically sound as supported by data in published human and animal studies.
The RfD is a risk assessment tool, not a risk management decision. Judgments
as to a "safe" dose and exposure are decisions that involve risk management
components.
Two new major prospective longitudinal studies, one in the Seychelles Islands and
the other in the Faroe Islands, have recently begun to publish their findings in the
literature. In November 1998, a federally sponsored workshop, Scientific Issues
Relevant to Assessment of Health Effects from Exposure to Methylmercury,
concluded that the results from the Faroe and Seychelles Islands studies are
credible and provide valuable new information on the potential health effects of
methylmercury. Significant uncertainties remain, however, because of issues
related to exposure, neurobehavorial endpoints, confounders and statistics, and
design (NIEHS, 1999).
The Science Advisory Board stated that the Seychelles and Faroe Island studies
have advantages over the studies in Iraq and New Zealand; they have much
larger sample sizes, a larger number of developmental endpoints, potentially more
sensitive developmental endpoints, and control a more extensive set of
potentially confounding factors. However, the studies also have some limitations
in terms of low exposures and ethnically homogeneous societies. The SAB
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concluded that the interim RfD may need to be reassessed in terms of the most
sensitive endpoints from these new studies. The National Academy of Sciences
(MAS) conducted an independent assessment of the interim RfD. They concluded
"On the basis of its evaluation, the committees' consensus is that the value of
EPA's current RfD for methylmercury, 0.1 ug/kg per day, is a scientifically
justifiable level for the protection of public health." However, the MAS
recommended that the Iraqi study no longer be used as the scientific basis for the
RfD. They recommended that the developmental neurotoxic effects of
methylmercury reported in the Faroe Islands study be used for the derivation of
the RfD (MAS, 2000a).
5.2.3.5 Reproductive and Developmental Toxicity—
Data are available on reproductive and developmental effects in rats, mice, guinea
pigs, hamsters, and monkeys. Convincing data from a number of human studies
i.e., Minamata Japan) also indicate that methylmercury causes subtle to severe
neurologic effects depending on dose and individual susceptibility. EPA considers
methylmercury to have sufficient human and animal data to be classified as a
developmental toxicant.
Methylmercury accumulates in body tissue; consequently, maternal exposure
occurring prior to pregnancy can contribute to the overall maternal body burden
and result in exposure to the developing fetus. In addition, infants may be
exposed to methlymercury through breast milk. Therefore, it is advisable to
reduce methylmercury exposure to women with childbearing potential to reduce
overall body burden.
5.2.3.6 Mutagenicity—
Methylmercury appears to be clastogenic but not to be a point mutagen; that is,
mercury causes chromosome damage but not small heritable changes in DMA.
EPA has classified methylmercury as being of high concern for potential human
germ cell mutagenicity. The absence of positive results in a heritable mutagenicity
assay keeps methylmercury from being included under the highest level of
concern. The data on mutagenicity were not sufficient, however, to permit
estimation of the amount of methylmercury that would cause a measurable
mutagenic effect in a human population.
5.2.3.7 Carcinogenicity—
Experimental animal data suggest that methylmercury may be tumorigenic in
animals. Chronic dietary exposures of mice to methylmercury resulted in
significant increases in the incidences of kidney tumors in males but not in
females. The tumors were seen only at toxic doses of methylmercury. Three
human studies have been identified that examined the relationship between
methylmercury exposure and cancer. There was persuasive evidence of
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increased carcinogenicity attributable to methylmercury exposure in any of these
studies. Interpretation of these studies was limited by poor study design and
incomplete descriptions of methodology and/or results. EPA has not calculated
quantitative carcinogenic risk values for methylmercury (IRIS, 1999). EPA has
found methylmercury to have inadequate data in humans and limited evidence in
animals and has classified it as a possible human carcinogen, Group C.
All of the carcinogenic effects were observed in the presence of profound damage
to the kidneys. Tumors may be formed as a consequence of repair in the
damaged organs. Evidence points to a mode of action for methylmercury
carcinogenicity that operates at high doses certain to produce other types of
toxicity in humans. Given the levels of exposure most likely to occur in the U.S.
population, even among consumers of large amounts offish, methylmercury is not
likely to present a carcinogenic risk.
5.2.3.8 Special Susceptibilities—
The developing fetus is at greater risk from methylmercury exposure than are
adults. Data on children exposed only after birth are insufficient to determine if
this group has increased susceptibility to central nervous system effects of
methlymercury. In addition, children are considered to be at increased risk of
methylmercury exposure by virtue of their greater food consumption (mg food/kg
body weight) by comparison to adult exposures. Additional risk from higher
mercury ingestion rates may also result from the apparently decreased ability of
children's bodies to eliminate mercury.
5.2.3.9 Interactive Effects-
Potassium dichromate andatrazine may increase the toxicity of mercury, although
these effects have been noted only with metallic and inorganic mercury. Ethanol
increases the toxicity of methylmercury in experimental animals. Vitamins D and
E, thiol compounds, selenium, copper, and possibly zinc are antagonistic to the
toxic effects of mercury (ATSDR, 1999).
5.2.3.10 Critical Data Gaps-
Additional data are needed on the exposure levels at which humans experience
subtle, but persistent, adverse neurological effects. Data on immunologic effects
and reproductive effects are not sufficient for evaluation of low-dose methyl-
mercury toxicity for these endpoints.
5.2.3.11 Summary of EPA Health Benchmarks-
Chronic Toxicity 1 x I0~4mg/kg-d
Carcinogenicity Group C (possible human carcinogen).
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5.2.3.12 Major Sources—
ATSDR (1999), IRIS (1999), Shubat (1993a), Stern (1993), U.S. EPA (1997d).
5.2.4 Selenium
5.2.4.1 Background-
Selenium is an element that occurs naturally in many areas and is produced
through industrial processes. It is an essential nutrient with a recommended
dietary allowance (RDA) of 55 ug/d (0.055 mg) for adult men and women. The
Tolerable Upper Intake Level for adults is set at 400 ug/d (0.4 mg/d) based on
selenosis as the adverse effect (MAS, 2000b). ATSDR has identified daily intake
at nontoxic levels of approximately 0.05 to 0.15 mg/d (ATSDR, 1996a; HSDB,
1993). This is approximately equivalent to 7 x 10~4to2x 10"3 mg/kg-d in a70-kg
individual.
Selenium plays a critical role in the antioxidant enzyme glutathione peroxidase.
Selenium deficiency has been associated with muscle degeneration in humans.
A serious form of this, congestive cardiomyopathy (Keshan disease), has been
studied in areas of China with low naturally occurring levels of selenium.
Selenium has also been shown to have a protective effect against chemically
induced cancers in laboratory animals (Robbins et al., 1989). Although selenium
is an essential nutrient, it is toxic at high exposure levels and is mutagenic in
some test systems (ATSDR, 1996a).
Definitive information concerning the chemical forms of selenium found in fish is
not available (U.S. EPA, 1993a). Due to the lack of information on chemical
forms, the toxicities of a variety of selenium forms are included in the discussion
below. In some parts of the United States, particularly in western states, soil
concentrations lead to selenium levels in plants that can cause human exposure
at potentially toxic levels (ATSDR, 1996a). This exposure should be considered
in evaluating the overall exposure to selenium and in developing fish consumption
advisories.
5.2.4.2 Pharmacokinetics—
Selenium contained in food is generally associated with proteins as organic
selenium compounds. It is easily absorbed by the body and accumulates
primarily in the liver and kidneys. It accumulates to a lesser extent in the blood,
lungs, heart, testes, and hair. Most of the selenium that enters the body is quicky
excreted in the urine, feces, and breath (ATSDR, 1996a).
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5.2.4.3 Acute Toxicity—
Signs of acute selenium poisoning include difficulty in walking; labored breathing;
cyanosis of the mucous membranes; congestion of the liver; endocarditis and
myocarditis; degeneration of the smooth musculature of the Gl tract, gall bladder
and bladder; and erosion of the long bones. Subacute selenosis (prolonged
exposure at relatively high doses) causes impaired vision, ataxia, disorientation,
and respiratory distress (IRIS, 1999). "Blind staggers" disease is a disease in
livestock that results from acute consumption of plants high in selenium. It is
characterized by impaired vision, aimless wandering behavior, reduced
consumption of food and water, and paralysis (ATSDR, 1996a).
5.2.4.4 Chronic Toxicity—
IRIS provides an RfD of 0.005 mg/kg-d for selenium and selenium compounds
based on a NOAEL of 0.015 mg/kg-d from a 1989 human epidemiological study
that found clinical selenosis at the LOAEL of 0.023 mg/kg-d. The NOAEL was
calculated from regression analysis of blood selenium levels and selenium intake.
An uncertainty factor of 3 rather than 10 was used for intraspecies variability.
EPA has medium confidence in the study on which the RfD was based due to
some possible interactions that were not fully explored. But because there are
many animal and epidemiologic studies that support the principal study, EPA has
high confidence in the database and, consequently, in the RfD (IRIS, 1999).
In epidemiological studies of populations exposed to high levels of selenium in
food and water, discoloration of the skin, loss of nails and hair, excessive tooth
decay and discoloration, garlic odor in the breath and urine, lack of mental
alertness, and listlessness were reported (IRIS, 1999). In high-selenium areas of
China, peripheral anesthesia and pain in the limbs have been reported.
Exaggerated tendon reflexes, convulsions, paralysis, and hemiplegia were
estimated to occur at a minimum chronic exposure of 0.053 mg/kg-d. A NOAEL
of 0.027 mg/kg-d was estimated (ATSDR, 1996a).
In animals, neurological dysfunction, respiratory distress, skin lesions with
alopecia, necrosis and loss of hooves, emaciation, and liver toxicity as indicated
by increases in serum transaminases and alkaline phosphatase have been seen
(IRIS, 1999). Cows with high, naturally occurring dietary exposures were found
to have irritation in the upper Gl tract (ATSDR, 1996a; IRIS, 1999).
Lifetime exposure of mice to sodium selenate or sodium selenite at 0.57 mg/kg-d
caused amyloidosis of the lung, liver, kidney, adrenal gland, and heart. Mice
appear to be more sensitive to selenium with regard to lung toxicity than rats.
(ATSDR, 1996a).
Hematological effects have been observed in multiple acute and chronic animal
studies. Rats subchronically exposed to wheat containing selenium at a dose of
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0.56 mg/kg-d for 6 weeks had a 79 percent reduction of blood hemoglobin
(ATSDR, 1996a).
Bone softening in rats has been noted with an LOAEL of 0.2 mg/kg-d with
exposure over several months (less than 100 days). Other musculoskeletal
effects have also been observed in livestock. Adverse effects on the liver and
kidneys have been observed in multiple animal studies with LOAELs of 0.1
mg/kg-d and above. Endocrine effects have been observed in animals fed
seleniferous wheat at doses of 0.4 mg/kg-d for 6 weeks. Dermal effects have
been observed at doses as low as 0.016 mg/kg-d in humans with dietary
exposure (ATSDR, 1996a). Depression of the immune system was observed in
rats exposed subchronically to sodium selenite at 0.7 mg/kg-d. At lower doses
(0.07 mg/kg-d and 0.28 mg/kg-d), mixed results were obtained, with a stimulation
of some components of the immune system and depression of others (ATSDR,
1996a).
5.2.4.5 Reproductive and Developmental Toxicity—
Limited information is available on the reproductive and developmental toxicity of
selenium in humans. In animals, selenium has caused growth retardation,
decreased fertility, embryotoxicity, fetotoxicity, and teratogenic effects.
A multigenerational study in mice dosed with selenate at 0.39 mg/kg-d identified
a significant increase in young deaths in the F1 generation and increased runts
in the F1 through F3 generations. Because only one dose was used, only a
LOAEL can be obtained from this study. A one-generation mouse study found a
NOAEL of 0.39 mg/kg-d. An early five-generation study identified a NOAEL of
0.075 mg/kg-d and a LOAEL of 0.125 mg/kg-d with a 50 percent reduction in the
number of young reared at that dose (IRIS, 1999).
Multiple studies have determined that exposure of livestock (e.g., sheep, pigs,
cattle) to naturally seleniferous diets resulted in fetal malformations and
interference with normal fetal development. Malformations were associated with
other manifestations of toxicity. The specific selenium compounds associated
with these effects have not been identified (ATSDR, 1996a). At 0.4 mg, pigs
exposed from 8 weeks of age had offspring with significantly reduced birth weight
and weaning weights (ATSDR, 1996a).
Chronic exposure studies in animals have identified multiple adverse effects on
the reproductive ability of animals and on offspring viability. Effects include:
altered menstrual cycles in monkeys exposed to 0.08 mg/kg-d for 30 days,
reduced rates of conception at 0.4 mg/kg-d in pigs exposed from 8 weeks of age
(other offspring effects are listed under developmental effects), abnormal length
estrus cycles in rats exposed subchronically to 0.31 mg/kg-d, increased fetal
resorption and decreased conception rate in livestock exposed at a LOAEL of
approximately 0.5 mg/kg-d, failure to breed in a three-generation study of mice
exposed at 0.57 mg/kg-d, no effects in a two-generation study of rats at 0.21
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mg/kg-d, and a 50 percent reduction in the number of young successfully reared
with maternal exposure to 0.35 mg/kg-d for 1 year. Male fertility also appears to
be affected by selenium exposure. Decreased sperm counts have been observed
in male rats exposed subchronically to 0.1 mg/kg-d and higher while abnormal
sperm and decreased testicular weights were observed at 0.2 mg/kg-d (ATSDR,
1996a).
5.2.4.6 Mutagenicity—
Data on the mutagenicity of selenium and its compounds are mixed. There are
many positive mutagenicity assays on selenium compounds including
unscheduled DMA synthesis, increased chromosomal aberrations in human
lymphocytes and in the bone marrow of rats, and an increase in sister chromatid
exchanges in human whole-blood cultures. There are also assays with negative
results (IRIS, 1999).
Inorganic selenium compounds appear to have genotoxic effects at relatively high
doses and antigenotoxic effects at lower doses. For example, a study of mice
exposed to mutagens and given doses of 0.05 to 0.125 mg/kg-d of selenium
indicates that selenium may inhibit the mutagenic effects of chemical agents
(ATSDR, 1996a).
5.2.4.7 Carcinogenicity—
Epidemiological studies that used the selenium concentration in crops as an
indicator of dietary selenium have generally reported an inverse association
between selenium levels and cancer occurrence. Animal studies have reported
that selenium supplementation results in a reduced incidence of several tumor
types (ATSDR, 1996; IRIS, 1999). EPA has determined that selenium is not
classifiable as to its carcinogenicity in humans (Group D) because of insufficient
data. EPA has classified selenium sulfide, an insoluble salt, as a probable human
carcinogen (B2) based on liver and lung tumors in oral exposure studies in
multiple species (IRIS, 1999).
5.2.4.8 Special Susceptibilities—
ATSDR has listed the following groups as potentially having greater susceptibility:
pregnant women and their fetuses, persons exposed to high fluoride levels in
drinking water (evidence equivocal), those with vitamin E deficiencies, and insulin-
dependent diabetics (ATSDR, 1996a).
5.2.4.9 Interactive Effects-
Selenium alters the toxicity of many chemicals. It reduces the toxicity of mercury,
cadmium, lead, silver, and copper. Most forms of selenium interact with arsenic
to reduce the toxicity of both elements. Selenium also interacts with vitamins,
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sulfur-containing amino acids, xenobiotics, and essential and nonessential
elements (ATSDR, 1996a).
5.2.4.10 Critical Data Gaps—
ATSDR has reported the following data gaps: human epidemiological data for all
relevant effects, relationship between selenium dietary exposure levels and
cancer; mechanisms of genotoxicity, reproductive, immunotoxicity, neurotoxicity,
especially behavioral and histopathological CMS effects, pharmacokinetic, and
bioaccumulation; and bioavailability from environmental media (ATSDR, 1996a).
5.2.4.11 Summary of EPA Health Benchmarks-
Chronic Toxicity 5 x 10~3 mg/kg-d
Carcinogenicity Group D (not classifiable).
5.2.4.12 Major Sources—
ATSDR (1996a), HSDB (1993), IRIS (1999).
5.2.5 Tributyltin Oxide
5.2.5.1 Background—
Tributyltin oxide belongs to the organometallic family of tin compounds that have
been used as biocides, disinfectants, and antifoulants. This compound and other
tributyltin compounds have high bioconcentration factors in aquatic organisms and
are acutely and chronically toxic to these organisms at low concentrations.
Because of concerns over these compounds' effects on nontarget aquatic
species, in 1986 EPA initiated a special review of tributyltin compounds used as
antifoulants (U.S. EPA, 1986e). In 1988, the Organotin Antifouling Paint Control
Act (OAPCA) was enacted, which contained interim and permanent tributyltin
restrictions as well as environmental monitoring, research, and reporting
requirements.
The tributyltin compounds registered for use as antifoulants are: tributyltin oxide,
tributyltin adipate, tributyltin dodecenyl succinate, tributyltin sulfide, tributyltin
acetate, tributyltin acrylate, tributyltin fluoride, tributyltin methacrylate, and
tributyltin resinate (U.S. EPA, 1986e). This toxicological profile discusses only
tributyltin oxide, since this is the only tributyltin compound with risk assessment
information (an RfD) and there is more toxicological information on this compound
than any other.
5.2.5.2 Pharmacokinetics—
The pharmacokinetic information available consists of data on organotin
compounds as a group; there are few data specific to tributyltin oxide. Organotin
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compounds appear to be absorbed in mammals, with studies in rats showing
detection of tin compounds in the gastrointestinal tract, kidney, and liver, with little
retention observed in the brain and blood. One study specific to tributyltin oxide
found the highest levels of tin in the liver and kidneys, with levels in the brain and
adipose tissue at 10 to 20 percent of the liver and kidney levels. The metabolism
of organotin compounds appears to involve dealkylation, with the liver as the
active site. There are no data regarding the excretion of organotin compounds
(ATSDR, 1992).
5.2.5.3 Acute Toxicity—
The limited available data show tributyltin oxide to be quite toxic to animals, with
oral LD50s ranging between 122 and 194 mg/kg in rats (ATSDR, 1992; HSDB,
1999) and 52 to 130 mg/kg in mice (WHO, 1999).
5.2.5.4 Chronic Toxicity—
There are no studies on the effects of tributyltin oxide in humans. Animal studies
have shown effects on the blood (lowered corpuscular volume and hemoglobin
mass and decreased leukocytes) and liver, and immunological effects including
thymus atrophy and depletion of T-lymphocytes in the spleen and lymph nodes
from tributyltin exposure (ATSDR, 1992; HSDB, 1999).
IRIS provides an RfD for tributyltin oxide of 3.0 x 10~4 mg/kg-d, based on a
benchmark dose (10 percent relative change as the benchmark response) of 0.03
mg/kg-d and an uncertainty factor of 100. This was based on a chronic rat
feeding study in which immunotoxicity was observed. The uncertainty factor of
100 reflects the uncertainty in extrapolating from laboratory animals to humans
and the uncertainty in the range of human sensitivity (IRIS, 1999; U.S. EPA,
1997g).
EPA has high confidence in the studies on which the RfD was based, medium to
high confidence in the overall database, and medium to high confidence in the
RfD. This is based on the fact that the principal study was a well-designed and
well-conducted chronic toxicity assay.(IRIS, 1999; U.S. EPA, 1997g).
5.2.5.5 Reproductive and Developmental Toxicity—
No studies are available on the reproductive and developmental effects of
tributyltin oxide in humans. In a two-generation reproductive study in rats, there
were no effects on mating, pregnancy, fertility, litter size, or pup survival in either
generation. Compound-related developmental effects were limited to decreased
pup body weight during lactation in both generations at the high dose. The
NOAEL for reproductive toxicity in this study was 4.4 mg/kg-d, the highest dose
tested. The NOAEL for developmental toxicity was 0.34 mg/kg-d (U.S. EPA,
1997g). When pregnant rats were exposed to high doses of tributyltin oxide (>10
mg/kg-d), decreased numbers of live births and decreased growth and viability of
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the offspring were reported. While these findings demonstrate the fetotoxic
potential of tributyltin oxide, a nonspecific effect of tributyltin oxide cannot be ruled
out because of overt maternal toxicity seen at the doses used (HSDB, 1993). A
developmental study in mice reported dose-related decreases in fetal weights,
some skeletal abnormalities, such as fused ribs and cleft palates, at all dose
levels and also in the controls. Weaknesses of this study include the occurrence
of developmental effects in both treated and control animals, maternal toxicity,
and lack of information on the statistical evaluation of the data (ATSDR, 1992;
U.S. EPA, 1997g).
5.2.5.6 Mutagenicity—
Results from in vitro studies on tributyltin oxide have been primarily negative.
Tributyltin oxide was negative in a variety of studies with Salmonella typhimurium
and Chinese hamster cells; the only positive results were with metabolic
activation. In vivo studies were also mainly negative; the compound was negative
in Drosophila melanogaster and in the micronucleus test (at cytotoxic doses) in
mice. One positive result was obtained in the micronucleus test where increased
micronuclei in erythrocytes were noted (ATSDR, 1992; HSDB, 1999).
5.2.5.7 Carcinogenicity—
No human studies are available. Cancer bioassays following oral exposure have
been conducted in rats and mice. The study in rats noted an increased incidence
of some benign tumors at the highest dose level. However, this study is
inconclusive because of increased mortality at the high dose and variable
background rates for the tumors observed. In the mouse study, no increase in
tumor incidence was observed. EPA has classified tributyltin oxide as Group D for
carcinogenicity - not classifiable as to human carcinogenicity (U.S. EPA, 1997g).
5.2.5.8 Special Susceptibilities-
There is some evidence that a child might be more sensitive to the toxic effects
of tributyltin oxide. For example, preweanling rats were shown to be more
sensitive than adult rats to the immunotoxic effects of tributyltin oxide. Because
the RfD is based on the effects observed when weanlings were dosed for the
remainder of their lives, any potential childhood sensitivity is already accounted
for. Animal toxicity studies showed no evidence of gender differences in the toxic
responses to tributyltin oxide (U.S. EPA, 1997g).
5.2.5.9 Interactive Effects-
Limited information is available on the interactive effects of tributyltin oxide.
Sulfur-containing compounds have been shown, in vitro, to interact with tributyltin
compounds to produce other compounds with lower hemolytic activity (ATSDR,
1992).
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5.2.5.10 Critical Data Gaps-
No human data are available to characterize the toxicity of tributyltin oxide. A
wealth of data from laboratory animals, however, is available. These data
adequately characterize the noncancer toxicity from oral exposure to tributyltin
oxide. EPA has high confidence in this assessment. The species studied include
monkey, dog, rat, and mouse. In addition, there is a two-generation reproduction
study and several developmental studies in rats and mice. The principal study and
a variety of supporting studies convincingly demonstrate that the critical effect for
tributyltin oxide is immunotoxicity. The potential for neurotoxicity has not been
completely studied (U.S. EPA, 1997g).
5.2.5.11 Summary of EPA Health Benchmarks-
Chronic Toxicity 3.0 x 10~4 mg/kg-d
Carcinogenicity Group D (not classifiable).
5.2.5.12 Major Sources—
ATSDR (1992a), HSDB (1999), IRIS (1999), U.S. EPA (1997g).
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5.3 ORGANOCHLORINE PESTICIDES
5.3 ORGANOCHLORINE PESTICIDES
5.3.1 Chlordane
5.3.1.1 Background—
Chlordane is an organochlorine insecticide comprised of the sum of cis- and trans-
chlordane and frans-nonachlor and oxychlordane for purposes of health advisory
development (U.S. EPA, 1997e). First introduced in 1947, it was used extensively
on agricultural crops, livestock, lawns, and for termite control. Because of
concern over cancer risk, human exposure, and effects on wildlife, most uses
were banned in 1978, and all uses were banned by 1988. Due to its long half-life
and ability to concentrate in biological materials, it is still widely distributed in fish
in the United States.
5.3.1.2 Pharmacokinetics—
Chlordane is extremely lipid soluble, and lipid partitioning of Chlordane and its
metabolites has been documented in both humans and animals. Concen trations
of chlordanes (cis- and trans-isomers and metabolites) detected in human liver
samples were 17-fold higher when expressed on a fat rather than a wet weight
basis. Chlordane is metabolized via oxidation, which results in a number of
metabolites, including oxychlordane, that are very persistent in body fat.
Reductive dehalogenation of Chlordane forms free radicals, which are
hypothesized to be significant in Chlordane toxicity (ATSDR, 1994a).
Human studies have found Chlordane in pesticide applicators, residents of homes
treated for termites, and those with no known exposures other than background
(e.g., food or airborne). Human milk fat contained a mean Chlordane residue of
approximately 188 ppm. Oxychlordane residues were detected in 68 percent of
human milk samples in a low-pesticide-usage area and in 100 percent of the 50
samples tested in Hawaii. It is anticipated that all routes of exposure were
involved in maternal exposure to Chlordane. Fat accumulation of Chlordane
appears to depend on the exposure duration (ATSDR, 1994a).
Mechanisms of toxicity include: the binding of Chlordane and its metabolites
irreversibly to cellular macromolecules, causing cell death or disrupting normal
cellular function; increasing tissue production of superoxide radicals, which
accelerates lipid peroxidation and disrupts the function of membranes; possible
suppression of hepatic mitochondrial energy metabolism; and alteration of
neurotransmitter levels in various regions of the brain; a reduction in bone marrow
stem cells prenatally; and suppression of gap junction intercellular communication
(ATSDR, 1994a).
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5.3.1.3 Acute Toxicity—
Chlordane is moderately to highly toxic with an estimated lethal dose to humans
of 6 to 60 g (I RIS, 1999). Effects reported in humans after acute exposure include
headaches, irritability, excitability, confusion, incoordination, seizures, and
convulsions. There is also some evidence that acute exposures to chlordane may
be associated with immunologic dysregulation, aplastic anemia in humans (U.S.
EPA, 1997e).
5.3.1.4 Chronic Toxicity—
IRIS provides an RfD of 5.0 x 10"4 mg/kg-d based on a NOAEL of 0.15 mg/kg-d
for hepatic necrosis in a 2-yr feeding study in mice (IRIS, 1999). The LOAEL in
the principal study was 0.75 mg/kg-d. An uncertainty factor of 300 was applied
to the NOAEL, 10 each for inter- and intraspecies variability and 3 for lack of any
reproductive studies. The confidence in the principal study is rated medium, as
is the confidence in the database.
Multiple neurological effects have been reported in humans exposed both acutely
and chronically to chlordane. When adults (109 women and 97 men) who had
been exposed to uncertain levels of chlordane via both air and oral routes were
examined, significant (p < 0.05) differences were observed with a battery of
neurophysiological and neuropsychological function tests (U.S. EPA, 1997e).
Also, profiles of mood states (including tension, depression, anger, vigor, fatigue,
and confusion) all were affected significantly (p < 0.0005) as compared to a
referent population.
5.3.1.5 Reproductive and Developmental Toxicity—
According to the IRIS file, "there have been 11 case reports of CMS effects, blood
dyscrasias and neuroblastomas in children with pre/postnatal exposure to
chlordane and heptachlor" (IRIS,1999).
ATSDR reports a number of developmental effects. Prenatal and early postnatal
exposure in mice may have permanent effects on the immune system, including
a reduction in the number of stem cells required to form the mature immune
system. Effects were observed at 4 mg/kg-d. Neurological effects include
abnormal behavior and increased seizure thresholds in mice at 1 mg/kg-d prenatal
and postnatal (via lactation) exposure (no NOEL was identified). Alterations in
plasma corticosterone levels were observed, which may result from a change in
the neuroendocrinological feedback mechanisms (ATSDR, 1994a).
Concerning cancer in children, see the discussion in Section 5.3.1.7.
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5.3.1.6 Mutagenicity—
Mutagenicity assays of chlordane have yielded mixed results, with positive results
generally obtained in higher organism cell assays and negative results in bacterial
assays (IRIS, 1999).
5.3.1.7 Carcinogenicity—
Chlordane is classified as a probable human carcinogen (B2) by EPA based on
oral studies in animals. An increased incidence of hepatocellular carcinoma was
observed in both sexes in mice in two separate studies using different strains.
Hepatocellular carcinomas were also observed in another study in male mice
using a third strain. The oral cancer slope factor of 0.35 per mg/(kg-d) is the
geometric mean of the cancer potencies calculated from five data sets (IRIS,
1999).
Five compounds structurally related to chlordane (aldrin, dieldrin, heptachlor,
heptachlor epoxide, and chlorendic acid) have produced liver tumors in mice.
Chlorendic acid also has produced liver tumors in rats.
Neuroblastoma and acute leukemia have also been associated with prenatal and
early childhood exposure to chlordane (ATSDR, 1994a).
5.3.1.8 Special Susceptibilities-
Based on the results of animal studies showing prenatal exposure causes
damage to the developing nervous and immune systems, fetuses and children
may be at greater risk than adults from chlordane exposure. According to
ATSDR:
Given the generally greater sensitivity to toxicants of incompletely
developed tissues, it seems possible that prenatal exposure of
humans to chlordane could result in compromised immunocom-
petence and subtle neurological effects (ATSDR, 1994a).
Due to the interactive effects of chlordane with other chemicals via microsomal
enzymes (see Section 5.3.1.9), ATSDR has cautioned that: "doses of therapeutic
drugs and hormones may require adjustment in patients exposed to chlordane."
The results of an acute animal study suggest that protein-deficient diets may also
increase the toxic effects of chlordane (ATSDR, 1994a).
ATSDR has listed the following populations as unusually susceptible: those with
liver disease or impaired liver function; infants, especially those with a hereditary
predisposition to seizures; and the fetus (ATSDR, 1994a).
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5.3.1.9 Interactive Effects—
Chlordane is a potent inducer of hepatic microsomal enzymes. Chlordane
exposure has been associated with an increased rate of metabolism of
therapeutic drugs, hormones, and many other endogenous and xenobiotic
compounds. Exposure to other chemicals that induce the same enzymes may
increase the toxicity of Chlordane by enhancing its metabolism to its toxic
intermediate. The acute toxic effects of aldrin, endrin, and methoxychlor with
Chlordane were greater than the additive sum of the individual toxicities (ATSDR,
1994a).
It has been suggested that increased dietary vitamins C or E or selenium may be
protective against free-radical-induced toxicity (ATSDR, 1994a).
MIXTOX reported synergistic effects between Chlordane and endrin in mice
exposed via gavage and both potentiation and inhibition with y-hexachloro-
cyclohexane in rodents exposed via gavage. Synergism is reported with
toxaphene and malathion together with Chlordane in mice exposed via gavage
(MIXTOX, 1992).
5.3.1.10 Critical Data Gaps—
IRIS lists the following data gaps for Chlordane: chronic dog feeding study, rat
reproduction study, rat teratology study, and rabbit teratology study (IRIS, 1999).
Other studies that are needed include a multigeneration study, which includes a
measurement of reproductive system toxicity, immunological effects—particularly
with developmental exposures, pharmacokinetic studies, and studies to determine
methods for reducing body burden (ATSDR, 1994a).
5.3.1.11 Summary of EPA Health Benchmarks-
Chronic Toxicity 5 x 10~4 mg/kg-d
Carcinogenicity 0.35 per mg/kg-d.
5.3.1.12 Major Sources—
ATSDR (1994a), HSDB (1993), IRIS (1999), EPA (1997e).
5.3.2 DDT, DDE, ODD
5.3.2.1 Background—
DDT is an organochlorine pesticide that has not been marketed in the United
States since 1972 but is ubiquitous due to its widespread use in previous decades
and its relatively long half-life. DDT's close structural analogs, DDE and ODD, are
metabolites of DDT and have also been formulated as pesticides in the past
(Hayes, 1982). DDT is very widely distributed; it has been found in seals in
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Finland and reptiles in the Everglades (HSDB, 1993). The NHANES II study
(National Human Monitoring Program of the EPA) detected DDE, a metabolite of
DDT, in 99 percent of the 12- to 74-yr-old study subjects (living in the Northeast,
Midwest, and South). The median level was 11.8 ppb in blood serum (HSDB,
1993).
Although some use of DDT continues throughout the tropics, it remains of human
health concern in the United States primarily due to its presence in water, soil,
and food (Hayes, 1982). Because individuals are typically exposed to a mixture
of DDE, DDT, and ODD and their degradation and metabolic products (ATSDR,
1994b), the sum of the 4,4'- and 2,4'- isomers of DDT, DDE, and ODD should be
considered in the development of fish consumption limits for this group of
chemicals (U.S. EPA, 1993a).
5.3.2.2 Pharmacokinetics—
DDT and its analogs are stored in fat, liver, kidney, and brain tissue; trace
amounts can be found in all tissues (Hayes, 1982). DDE is stored more readily
than DDT (Hayes, 1982). DDT is eliminated through first-order reduction to ODD
and, to a lesser extent, to DDE. The ODD is converted to more water-soluble bis
(p-chlorophenyl)-acetic acid, with a biological half-life of 1 year. DDE is eliminated
much more slowly, with a biological half-life of 8 years. Because elimination
occurs slowly, ongoing exposure may lead to an increase in the body burden over
time.
5.3.2.3 Acute Toxicity—
The low effect dose for severe effects (acute pulmonary edema) in infants has
been reported to be 150 mg/kg. In adults, behavioral effects were noted at 5 to
6 mg/kg and seizures at 16 mg/kg (HSDB, 1993).
Evidence from acute exposure studies of dogs indicates that DDT may sensitize
the myocardium to epinephrine. This was observed for both injected epinephrine
and epinephrine released by the adrenal glands during a seizure and resulted in
ventricular fibrillation (Hayes, 1982). DDT may concurrently act on the CNS, in
a manner similar to that of other halogenated hydrocarbons, to increase the
likelihood of fibrillation (Hayes, 1982). Chronic exposure to 10 mg/kg-d did not
produce increased incidence of arrhythmias in rats or rabbits (Hayes, 1982).
ODD is considered less toxic than DDT in animals. Symptoms develop more
slowly and have a longer duration with ODD than with DDT exposure. Lethargy
is more significant and convulsions are less common than with DDT exposure
(HSDB, 1993).
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5.3.2.4 Chronic Toxicity—
Extensive research has been conducted on chronic and subchronic exposure
effects of DDT in animals and in humans working with DDT. These studies have
primarily focused on carcinogenic effects, which are discussed in Section 5.3.2.7.
Studies have also identified liver damage, and there is limited evidence that DDT
may cause leukocytosis and decreased hemoglobin level (Hayes, 1982).
Immunological effects have been associated with exposure to DDT. Exposure to
DDT at 2.63 mg/kg-d for 10 days resulted in immunological effects in rabbits.
With 31 days of exposure at 1 mg/kg-d in rats, a decrease in the number of mast
cells was observed. A relatively recent 8-week study in rabbits found decreases
in germinal centers of the spleen and atrophy of the thymus at 0.18 mg/kg-d.
Other effects were observed at higher doses. No studies were provided on
immunological effects following chronic exposure (ATSDR, 1994b).
IRIS lists an oral RfD of 5 x 10~4 mg/kg-d for DDT based on liver effects with a
NOAEL of 0.05 mg/kg-d from a 27-wk rat feeding study conducted in 1950.
Uncertainty factors of 10 each for inter- and intraspecies variability were used;
however, the usual factor of 10 for a less-than-lifetime study was not applied
"because of the corroborating chronic study in the data base" (IRIS, 1999). The
corroborating study was conducted in 1948.
5.3.2.5 Reproductive and Developmental Toxicity—
DDT causes embryotoxicity and fetotoxicity but notteratogenicity in experimental
animals (ATSDR, 1994b). Studies indicate that estrogen-like effects on the
developing reproductive system occur (ATSDR, 1994b). This also occurs with
chronic exposure as discussed in Section 5.3.2.4. Rabbits exposed to 1 mg/kg-d
early in gestation had decreased fetal brain, kidney, and body weights (ATSDR,
1994b). Prenatal exposure in mice at 1 mg/kg on 3 intermittent days resulted in
abnormal gonad development and decreased fertility in offspring, which was
especially evident in females (Hayes, 1982).
A three-generation rat reproduction study found increased offspring mortality at
all dose levels with a LOAEL of 0.2 mg/kg-d. Three other reproduction studies
found no effects at much higher dose levels (IRIS, 1999). Effects on the
urogenital system were found with 8 days' prenatal exposure in mice. Behavioral
effects in mice exposed prenatally for 7 days were noted at 17.5 mg/kg-d (HSDB,
1993).
Prenatal 1-day exposure of rabbits to DDT resulted in an abnormal persistence
of preimplantation proteins in the yolk sac fluid. The results suggest that DDT
caused a cessation of growth and development before implantation or during later
uterine development. The authors suggest that damage can be repaired but may
result in offspring with prenatal growth retardation in the absence of gross
abnormalities (HSDB, 1993). Most dosages tested for these effects have been
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relatively high. Postnatal exposure of rats for 21 days to 21 mg/kg (the only dose
tested) resulted in adverse effects on lactation and growth.
In dogs, placental passage of DDT to the fetus has been demonstrated. This was
confirmed in mice. Primary targets include the liver, adipose tissue, and intestine.
Rabbit blastocysts (a very early stage of development) contained a significant
amount of DDT shortly after administration to the mother (HSDB, 1993).
Biomagnification in human milk has been observed. In lactating women with an
intake of 5 x 10~4 mg/kg-d of DDT, the milk contained 0.08 ppm. This was
calculated to result in infant doses of 0.0112 mg/kg-d, which is approximately 20
times the dosage to the mothers (HSDB, 1993).
DDT is suspected of causing spontaneous abortion in humans and cattle (Hayes,
1982). The average concentration of DDE in the blood of premature babies
(weighing <2,500 g) was significantly greater than those of higher birth weight
infants (HSDB, 1993). The relationship between spontaneous abortion,
premature delivery, and maternal exposure and body burden requires clarification.
DDT accumulates in body tissue; consequently, exposure occurring prior to
pregnancy can contribute to the overall maternal body burden and result in
exposure to the developing individual. As a result, it is necessary to reduce
exposure to children and females with childbearing potential to reduce overall
body burden. If a female has been exposed to DDT, even if exposure is reduced
during pregnancy, the outcome of that pregnancy may be affected, depending on
the timing and extent of prior exposure.
DDT may have reproductive system toxicity. It appears to bind to uterine tissue
and have estrogenic activity (Hayes, 1982). Metabolites of DDT bind to the
cytoplasmic receptor for estrogen, which may result in inadvertent hormonal
response (agonist) or depress normal hormonal balance (antagonist). Either may
result in reproductive abnormalities (HSDB, 1993). The animal studies of the
reproductive system have yielded mixed results. Chronic animal studies have
identified LOELs that range over orders of magnitude. Serious adverse effects
(decreased fertility and decreased litter size) have been observed at 0.35 and
0.91 mg/kg-d, respectively, in subchronic animal studies. Edema of the testes
occurred at 2 mg/kg-d in a rat study. NOAELs are not available for these studies.
Other studies have identified NOAELs ranging from 2.4 to 10 mg/kg-d with severe
effects at 12 mg/kg-d (increased maternal and offspring death) (ATSDR, 1994b).
Significant reproductive (function and lactation) abnormalities have also been
observed at higher doses (83 mg/kg-d in rats and at 33.2 mg/kg-d in mice).
Function abnormalities have also been observed in dogs (Hayes, 1982).
5.3.2.6 Mutagenicity—
Genotoxicity studies in human systems strongly suggest that DDT may cause
chromosomal damage (ATSDR, 1994b). This is supported by in vitro and in vivo
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studies in animals (ATSDR, 1994b) and in some bacterial assays (HSDB, 1993).
There are multiple positive assays including human lymphocytes, human
leukocytes, human fibroblasts, an oncogenic transformation, and unscheduled
DMA synthesis in rats in multiple studies (ATSDR 1994b; HSDB, 1993).
5.3.2.7 Carcinogenicity—
DDE, DDT, and ODD are all considered probable human carcinogens (B2) based
on animal studies, with cancer potencies of 0.24, 0.34, and 0.34 per mg/kg-d,
respectively (IRIS, 1999). Liver tumors were associated with each chemical. It
is noted in the IRIS file that 24 of the 25 carcinogenicity assays of DDT have
yielded positive results. The occupational studies of workers exposed to DDT are
of insufficient duration to assess carcinogenicity (IRIS, 1999). Elevated leukemia
incidence, particularly chronic lymphocytic leukemia, was noted in two studies of
workers. Lung cancer has also been implicated in one study. Bone marrow cells
in experimental animals have also been affected by exposure, including an
increase in chromosomal fragments in the cells (HSDB, 1993).
It is recommended that the total concentration of the 2,4'- and 4,4'-isomer of DDT
and its metabolites, DDE and ODD, be evaluated as a group using the cancer
potency of 0.34 per mg/kg-d (U.S. EPA, 1993a). In addition, the EPA Carcino-
genicity Assessment Group has recommended that this value be used for
combinations of dicofol with the above three compounds (U.S. EPA, 1993a).
5.3.2.8 Special Susceptibilities-
Based on the information obtained from a recent developmental study that found
neurotoxicity and structural brain alterations at relatively low exposures
(approximately 50-fold less than in adults), children may be at greater risk from
DDT exposure than adults.
The results of the cardiac toxicity studies are not consistent; however, it is safest
to assume that exposure to DDT or its analogs may pose a risk for individuals
with cardiac disease at exposure levels estimated to be safe for the general
population (Hayes, 1982).
Individuals exposed to DDT may metabolize some drugs more rapidly than the
general population (HSDB, 1993). For example, increased phenobarbital
metabolism resulting from an increased body burden of DDT (10 ug) led to a 25
percent decrease in effectiveness of the drug in experimental animals. The
toxicity of chloroform was enhanced by the addition of DDT to the diet due to its
capacity as a microsomal stimulator (HSDB, 1993). Alterations in the metabolism
of drugs, xenobiotics, and steroid hormones may result from DDT exposure due
to DDT's induction of the hepatic mixed-function oxidase system at relatively low
doses (HSDB, 1993). Individuals who use medications that involve the mixed
function oxidase system directly (MFO inhibitors) or through metabolic processes
may be at risk for alteration of the drug's efficacy and/or timing if they are exposed
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to DDT. Information is not available for this document on the specific
relationships between various Pharmaceuticals and DDT/DDE/DDD body burdens
or intakes. This type of information merits further investigation.
ATSDR notes that persons with diseases of the nervous system or liver may be
particularly susceptible to the effects of DDT (ATSDR, 1994b). Based on
information discussed above concerning biomagnification in milk, nursing infants
may also be at greater risk due to their increased exposure.
5.3.2.9 Interactive Effects—
As discussed in Section 5.3.2.8, DDT exposure may alter the response to drugs,
xenobiotics, and endogenous steroid hormones. DDT is reported to promote
some tumorigenic agents and antagonize others. The actions may be related to
the induction of microsomal enzymes (ATSDR, 1994b).
5.3.2.10 Critical Data Gaps—
IRIS notes the lack of a NOAEL for reproductive effects and a relatively short
duration for the critical study on which the RfD is based.
Information was not located for this document on the specific relationships
between various Pharmaceuticals and DDT/DDE/DDD body burdens or intakes.
Information on the relationship between pre- and postnatal exposure and
behavioral effects and maternal exposure and milk concentrations is also needed.
An interagency group of researchers from NTP, ATSDR, and EPA have identified
the following data gaps: pharmacokinetic data; animal studies on respiratory,
cardiovascular, Gl, hematological, musculoskeletal, and dermal/ocular effects; the
significance of subtle biochemical changes such as the induction of microsomal
enzymes in the liver and the decreases in biogenic amines in the nervous system
in humans; an epidemiological study in humans of estrogen-sensitive cancers
including endometrial, ovarian, uterine, and breast cancer; reproductive system
toxicity; developmental toxicity; a multiple assay battery for immunotoxicity; subtle
neurological effects in humans; and mechanisms of neurotoxicity in the neonate
(ATSDR, 1994b).
5.3.2.11 Summary of EPA Health Benchmarks-
Chronic Toxicity 5 x 10~4 mg/kg-d (DDT only)
Carcinogenicity 0.34 per mg/kg-d. (sum of the 4,4' and 2,4'-isomers of DDT
DDE, and ODD)
5.3.2.12 Major Sources—
ATSDR (1994b), Hayes (1982), HSDB (1993), IRIS (1999).
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5.3 ORGANOCHLORINE PESTICIDES
5.3.3 Dicofol (Kelthane)
5.3.3.1 Background—
Dicofol is an organochlorine miticide/pesticide first registered for use in 1957.
Dicofol is used mainly on cotton, apples, and citrus crops; most of the use is in
California and Florida (U.S. EPA, 1998a). Dicofol is considered a DDT analog
based on its structure and activity (Hayes and Laws, 1991). In the past, dicofol
often contained 9 to 15 percent DDT and its analogs. In 1989, EPA required that
these contaminants constitute less than 0.1 percent of dicofol (HSDB, 1993).
5.3.3.2 Pharmacokinetics—
Studies with radiolabeled dicofol in rats indicated that most of the label was
eliminated in the feces after oral dosing (U.S. EPA, 1998a). Intact dicofol was
preferentially stored in adipose tissue. The major metabolic pathway was
reductive halogenation to dichlorodicofol and subsequent oxidation to more water-
soluble compounds.
5.3.3.3 Acute Toxicity—
The acute oral LD50 for dicofol in rats was 587 mg/kg (U.S. EPA, 1998a). A single
large oral dose of dicofol to rats caused ataxia at 350 mg/kg and weight loss at 75
mg/kg. The NOAEL for neurotoxicity in this study was 15 mg/kg. An acute dietary
RfD of 0.05 mg/kg-d was calculated based on this NOAEL and using an
uncertainty factor of 300 (U.S. EPA, 1998a).
5.3.3.4 Chronic Toxicity—
No RfD is currently listed in the IRIS file for this chemical (IRIS, 1999). The OPP
has recently derived an RfD of 0.0004 mg/kg-d for chronic dietary exposure (U.S.
EPA, 1998a). The critical effect was hormonal toxicity, based on inhibition of
adrenocortical trophic hormone (ACTH)-stimulated release of cortisol in dogs.
The NOAEL of 0.12 mg/kg-d was divided by an uncertainty factor of 300(1 OX for
interspecies variation, 10X for intraspecies extrapolation, and 3X for the protection
of infants and children.
5.3.3.5 Reproductive and Developmental Toxicity—
In a two-generation reproduction study in rats, the NOAEL for reproductive toxicity
was 0.4 mg/kg-d based on the ovarian vacuolation in the F1 females, an effect on
reproductive physiology. For offspring toxicity, the NOAEL was 2 mg/kg-d based
on decreased F2 pup viability (U.S. EPA, 1998a).
In a special one-generation postnatal toxicity study in rats, the NOAEL for both
offspring and parental toxicity was 1.7 mg/kg-d, based on histopathologic findings
in the liver. No treatment-related effects were observed on parameters of
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reproductive function or performance. The NOAEL for reproductive toxicity was
>9.8 mg/kg-d (U.S. EPA, 1998a).
No developmental toxicity was seen in a study in rats. The NOAEL was 25
mg/kg-d, the highest dose tested. In a developmental toxicity study in rabbits, the
NOAEL was 4 mg/kg-d, based on an increased incidence of abortions in the does
at 40 mg/kg-d (U.S. EPA, 1998a).
5.3.3.6 Mutagenicity—
Dicofol was negative for mutagenicity in the Ames test and for structural
chromosomal aberrations in Chinese hamster ovary cells. Dicofol did not induce
a clastogenic response in the chromosomes of rat bone marrow cells after oral
dosing (U.S. EPA, 1998a). Studies of dicofol in human lymphoid cells in vitro
were positive with an incidence of events 13 times that of controls. It induced
sister chromatid exchange with activation. Other mutagenicity studies in bacteria
have yielded negative results (HSDB, 1993).
5.3.3.7 Carcinogenicity—
In 2-yr carcinogenicity studies in mice and rats, dicofol administration resulted in
an increase in liver adenomas and combined liver adenomas and carcinomas in
male mice (U.S. EPA, 1998a). No increase in tumor incidence was observed in
female mice or in rats or in another 2-yr feeding study in either sex of rats. Dicofol
has been classified as a group C carcinogen (possible human carcinogen) based
on the increase in liver adenomas and combined liver adenomas and carcinomas
in male mice (U.S. EPA, 1998a).
5.3.3.8 Special Susceptibilities—
Toxicity data for dicofol provide no indication of increased susceptibility of rats or
rabbit fetuses following in utero exposures in the prenatal developmental toxicity
studies or following postnatal exposure in the two-generation reproduction study.
For this reason, the additional 10X Safety Factor for the protection of infants and
children was reduced to 3X (U.S. EPA, 1998a).
5.3.3.9 Interactive Effects—
As with other organochlorine pesticides, microsomal enzyme induction occurs and
may cause interactions with other chemicals. No additional data were located
(U.S. EPA, 1998a).
5.3.3.10 Critical Data Gaps—
EPA is requiring a developmental neurotoxicity study in rats for dicofol (U.S. EPA,
1998a). No other data gaps were identified (U.S. EPA, 1998a).
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5.3 ORGANOCHLORINE PESTICIDES
5.3.3.11 Summary of EPA Health Benchmarks-
Chronic Toxicity 4.0 x 10"4 mg/kg-d
Carcinogenicity Group C (possible human carcinogen).
5.3.3.12 Major Sources—
HSDB (1993), U.S. EPA (1993e), U.S. EPA (1998a).
5.3.4 Dieldrin
5.3.4.1 Background—
Dieldrin is an organochlorine pesticide that was phased out between 1974 and
1987. Dieldrin was mainly used on soil-dwelling pests and for termite control. It
continues to be detected nationwide due to its relatively long half-life. Dieldrin is
also a product of aldrin metabolism, a structurally similar organochlorine pesticide
which is also no longer in use (ATSDR, 1991).
5.3.4.2 Pharmacokinetics—
Dieldrin is absorbed from the Gl tract and transported via the hepatic portal vein
and the lymphatic system. It is found shortly after exposure in the liver, blood,
stomach, and duodenum. Dieldrin is lipophilic and is ultimately stored primarily
in fat and tissues with lipid components (e.g., brain) (ATSDR, 1991).
In dosing studies with volunteers at 0.0001 to 0.003 mg/kg-d over 2 years, the
time to achieve equilibrium was approximately 15 months. A dynamic equilibrium
was theorized with the average ratio of the concentration in adipose tissue to
blood of 156. Cessation of dosing led to decreases in blood levels following first-
order kinetics with a half-life ranging from 141 to 592 days and an average of 369
days (ATSDR, 1991).
The metabolism of dieldrin is described in detail in ATSDR (1991). Sex and
species differences have been reported in the metabolism and tissue distribution
of dieldrin based on chronic exposure studies and toxicokinetic studies in animals.
Males appear to metabolize and excrete dieldrin more rapidly than females
(ATSDR, 1991).
A correlation between exposure and dieldrin levels in human breast milk has been
established. Placental transfer of dieldrin has been observed in women, with
higher concentrations measured in fetal blood than in maternal blood (ATSDR,
1991).
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5.3.4.3 Acute Toxicity—
Acute effects include possible hematological effects in humans (pancytopenia and
thrombocytopenia, immunohemolytic anemia) (ATSDR, 1991). An estimated
human lethal dose is 65 mg/kg (HSDB, 1993).
5.3.4.4 Chronic Toxicity—
IRIS provides an RfD of 5 x 10'5 mg/kg-d based on a NOAEL of 0.005 mg/kg-d
from a 1969 2-year rat feeding study that found liver lesions (focal proliferation
and hyperplasia). Uncertainty factors of 10 each for inter- and intraspecies
variability were applied (IRIS, 1999). Liver toxicity has been observed in multiple
animal studies and in human acute exposure episodes. Adaptive changes (e.g.,
liver enlargement) have been observed at 0.00035 mg/kg-d in a subchronic rat
study.
Although the critical effect in the IRIS study was liver lesions, it was noted that, at
the next highest dose (0.05 mg/kg-d), "all animals became irritable and exhibited
tremors and occasional convulsions" (IRIS, 1999). There was no listing of
additional neurobehavioral studies in the IRIS file. As an organochlorine
pesticide, it is expected that dieldrin is a CMS toxicant. This is supported by acute
toxicity effects of dieldrin and the neurotoxicity studies listed below.
Other effects associated with dieldrin exposure include: arterial degeneration in
rats with a chronic exposure to 0.016 mg/kg-d, hematological disorders in
experimental animals at 0.25 and 1 mg/kg-d, musculoskeletal pathology at 0.015
mg/kg-d in a chronic rat study, kidney degeneration and other changes at 0.125
mg/kg-d in chronic animal studies in multiple species, hypertension in humans
(exposure level unknown), and multiple deficits in immune system function in
multiple studies (ATSDR, 1991). Increased susceptibility to tumor cells was
observed in a subchronic mouse study (dose not specified in material reviewed)
(HSDB, 1993).
Neurological effects of dieldrin have been observed in experimental animals and
in humans exposed acutely and chronically. Wheat mixed with aldrin and lindane
was consumed for 6 to 12 months by a small human population. Effects were
attributed to aldrin (converted to dieldrin via metabolism) because the wheat had
been mixed with lindane in previous years without adverse effect. A variety of
CMS disorders were observed, and abnormal EEGs were noted. Some
symptoms (myoclonic jerks, memory loss, irritability) continued for at least 1 year
after cessation of exposure. A child is believed to have developed mild mental
retardation as a result of exposure. Quantitative exposure information was not
available in the data reviewed (ATSDR, 1991).
Neurotoxicity has been observed in humans with chronic inhalation and dermal
exposures (ATSDR, 1991). Chronic exposure of pesticide applicators to dieldrin
led to idiopathic epilepsy, which ceased when exposure was terminated (HSDB,
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1993). Dermal and inhalation exposure were the likely routes of exposure. No
exposure quantitation was available.
A 1967 study of human exposure effects over 18 months at levels up to 0.003
mg/kg-d identified no effects on the CMS (as measured by EEG), peripheral nerve
activity, or muscle activity (ATSDR, 1991).
Animal studies have identified neurological effects including behavioral disorders
and learning deficits at doses of 0.1 to 0.25 mg/kg-d in subchronic and chronic
studies. Higher doses produced more dramatic effects (e.g., convulsions,
tremors). Cerebral edema and degeneration were found with chronic exposure
of rats to 0.016 mg/kg-d (ATSDR, 1991). Neural lesions (cerebral, cerebellar,
brainstem, and vascular) were observed in chronically exposed rats at 0.004
mg/kg-d (HSDB, 1993).
5.3.4.5 Reproductive and Developmental Toxicity—
IRIS provides limited information regarding the developmental toxicity of dieldrin.
A NOAEL of 6 mg/kg-d was obtained from a mouse teratology study with
exposure occurring from the 7th to 16th day of gestation. Fetotoxicity (decreased
numbers of caudal ossification centers and an increased incidence of extra ribs)
was observed with an LOAEL of 6 mg/kg-d. This study was not considered in
development of the IRIS file because 41 percent of the maternal fatalities
occurred at the LOAEL dose (IRIS, 1999).
A variety of effects in multiple organ systems have been observed in experimental
animals exposed prenatally to dieldrin. Skeletal anomalies and malformations
(e.g., cleft palate, webbed foot, open eyes, extra ribs) were identified at relatively
large doses (LEL of 3 mg/kg-d) (ATSDR, 1991).
Abnormalities of the CMS, eye, and ear were noted with a TD L0 (similar to a
LOAEL) of 30.6 mg/kg prenatal exposure, and craniofacial abnormalities were
observed at a single prenatal dose of 15 mg/kg-d (HSDB, 1993). Liver damage
has been observed in experimental animals at dosages as low as 0.016 mg/kg-d
(ATSDR, 1991). Note that liver lesions are the basis for the chronic toxicity RfD
derived from a study of adult animals, as reported in IRIS (IRIS, 1999). A
multigeneration study in mice found histological changes in liver, kidney, lungs,
and brain tissues in the first and second generation offspring at an LOAEL of 3
ppm (0.075 mg/kg-d) (HSDB, 1993).
Multiple studies have reported increased postnatal mortality following prenatal
exposure to dieldrin. Studies in dogs, rats, and mice have found LELs of 0.125
to 0.65 mg/kg-d associated with high mortality in offspring in the absence of
increased maternal mortality. Studies designed to evaluate the underlying causes
of mortality suggest that cardiac glycogen depletion, leading to cardiac failure,
may be causal (ATSDR, 1991).
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Neural lesions in prenatally exposed rats were found at an LOAEL of 0.004
mg/kg-d. Effects included cerebral edema, internal and external hydrocephalus,
and focal neuronal degeneration. Postnatal exposure of rats from day 5 of
gestation to 70 days of age resulted in increased learning ability at 3.5 x 10~4
mg/kg-d (the only dose tested). ATSDR has cautioned that "interpretation of the
results is difficult because the significance of improved performance in behavioral
paradigms is unknown, and the study is limited because only one dose of dieldrin
was tested" (ATSDR, 1991). In a rat multigeneration study, a TD L0 of 0.014
mg/kg-d with behavioral effects was observed (HSDB, 1993).
Dieldrin is known to accumulate in human milk. In one study of 102 samples in
the United States, 91.2 percent of the samples contained measurable levels of
dieldrin, with a mean concentration of 0.062 ppm lipid basis. Another U.S. study
found 80 percent of the 1,436 samples were positive with a range of 0.16 to 0.44
ppm milk fat (HSDB, 1993). This indicates that lactation may provide a significant
dietary source in infants with mothers who have been exposed to dieldrin. As
discussed above, studies in humans also determined that dieldrin can pass
through the placenta and is found in fetal blood.
Neurotoxicity appears to be a relatively sensitive endpoint for developmental
toxicity. The association of neurotoxic effects with dieldrin exposure is supported
by the observation of neurological effects in human populations exposed to
dieldrin. The study noted in the paragraph above that identified neural lesions
associated with prenatal exposure provided an LOAEL of 0.004 mg/kg-d provides
the most sensitive developmental toxicity measure of those reviewed. If the
LOAEL from this study were used to calculate an estimated exposure limit for
developmental effects, the standard uncertainty factors would typically take into
consideration inter- and intraspecies variability and the use of an LOAEL rather
than a NOAEL
As with the other organochlorines, it is anticipated that dieldrin can accumulate in
body tissue; consequently, exposure occurring prior to pregnancy can contribute
to the overall maternal body burden and result in exposure to the developing
individual. As a result, it is necessary to reduce exposure to children and females
with childbearing potential to reduce overall body burden. If a female has been
exposed to dieldrin, even if exposure is reduced during pregnancy, the outcome
of that pregnancy may be affected, depending on the timing and extent of prior
exposure.
Dieldrin causes reproductive system disorders in animals and one study suggests
that it may cause adverse effects in humans. In a study evaluating the blood and
placental levels of organochlorines associated with premature labor or
spontaneous abortions in women, positive results were obtained for aldrin. Most
exposed subjects had multiple chemical exposures; consequently, interpretation
of study results is difficult (ATSDR, 1991). See also notes regarding estrogenic
activity in Section 5.3.4.7.
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Studies of reproductive effects in animals indicate that exposure to dieldrin may
cause a number of adverse effects. Dieldrin exposure causes changes in the
levels of serum luteinizing hormone (LH) in females and gonadotropin in males.
Dieldrin interferes with the binding of dihydrotestosterone to male sex hormone
receptors (HSDB, 1993). These three hormones are critical to normal
reproductive function. A mouse study found decreased fertility with exposure to
1.3 mg/kg-d in females and 0.5 mg/kg-d in males. Another study found no effects
at much higher exposure levels. Adverse reproductive effects in dogs exposed
at LOAEL of 0.15 mg/kg-d for 14 months prior to mating included increased
stillbirth rates, delayed estrus, reduced libido, and a lack of mammary function
and development. Maternal behavior was studied in mice exposed for 4 weeks
prior to delivery until weaning at 1.95 mg/kg-d. Exposed maternal animals
violently shook the pups, ultimately killing them; others neglected their litters
(ATSDR, 1991).
5.3.4.6 Mutagenicity—
There is limited information on the mutagenicity of dieldrin. Positive in vivo
studies have found an increased incidence in the number of abnormal
metaphases in dividing spermatocytes and in univalents. Dominant lethal assays
(in vivo) have yielded mixed results. In vitro assays have also yielded mixed
results. Positive results have been obtained in cultured human lung cells and
mouse bone marrow cells (both found increases in chromosome aberrations) and
sister chromatid exchange (SCE) assays.
Dieldrin may not act directly on DMA; however, it may act by depressing transfer
RNA activity, increasing unscheduled DMA synthesis, and inhibiting metabolic
cooperation and gap junctional intercellular communication, according to
mechanistic studies. The inhibition of gap junctional communication may be
responsible for carcinogenic activity through depressing the cells' ability to control
excess proliferation. This inhibition has been correlated with strains and species
in which dieldrin has been shown to be carcinogenic. This type of activity is
considered promotion rather than initiation of tumors (ATSDR, 1991).
5.3.4.7 Carcinogenicity—
Dieldrin is classified as a probable human carcinogen (B2) by EPA based on oral
studies in animals. The oral cancer slope factor is 16 per mg/kg-d. Liver
carcinoma was identified in the animal studies. The geometric mean of 13 data
sets (with a range of a factor of 8) was used to develop the cancer potency (IRIS,
1999).
A variety of tumor types have been observed in animal studies including
pulmonary, lymphoid, thyroid, and adrenal (ATSDR, 1991). ATSDR has
concluded that dieldrin is probably a tumor promotor, based on genotoxicity and
mechanistic studies reviewed (ATSDR, 1991). Dieldrin has recently been
observed to have estrogenic effects on human breast cancer estrogen-sensitive
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cells (Soto et al., 1994). Xenoestrogens have been hypothesized to have a role
in human breast cancer (Davis et al., 1993). In addition to potential carcinogenic
effects, dieldrin may also cause disruption of the endocrine system due to its
estrogenic activity (Soto et al., 1994).
5.3.4.8 Special Susceptibilities—
ATSDR has identified the following populations as unusually susceptible: very
young children with immature hepatic detoxification systems, persons with
impaired liver function, and persons with impaired immune function (ATSDR,
1991). Based on the toxicity data reviewed above, individuals with the following
diseases or disorders may also be at increased risk: hypertension, hematological
disorders, musculoskeletal diseases, neurological diseases, and kidney disease.
The data also indicate that prenatal exposure may generate risks to children at
relatively low levels of exposure. Postnatal exposure, especially via lactation, may
also be a significant concern.
5.3.4.9 Interactive Effects—
In cows, dieldrin exposure increased the toxicity of diazinon; greater depression
in blood cholinesterase activity occurred, leading to severe clinical signs (HSDB,
1993).
MIXTOX has reported inhibition between dieldrin and hexachlorobenzene in rats
exposed orally via food. Studies have also reported additive effects (MIXTOX,
1992).
5.3.4.10 Critical Data Gaps—
A joint team of scientists from EPA, NTP, and ATSDR have identified the following
study data gaps: mechanism of animal carcinogenicity, genotoxicity in vivo and
in vitro, reproductive system toxicity, developmental toxicity, especially
mechanisms of postnatal mortality and teratogenesis, immunotoxicity,
neurotoxicity focusing on sensitive endpoints, and pharmacokinetics (ATSDR,
1991).
5.3.4.11 Summary of EPA Health Benchmarks-
Chronic Toxicity 5 x 10~5 mg/kg-d
Carcinogenicity 16 per mg/kg-d.
5.3.4.12 Major Sources—
ATSDR (1991), HSDB (1993), IRIS (1999).
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5.3.5 Endosulfan I, II
5.3.5.1 Background—
Endosulfan is an organochlorine pesticide comprised of stereoisomers designated
I and II, which have similar toxicities (U.S. EPA, 1993a). Endosulfan I and II are
referred to collectively as endosulfan; discussions refer to both isomers unless
otherwise noted. Endosulfan has been in use since 1954.
5.3.5.2 Pharmacokinetics—
Endosulfan is absorbed through the Gl tract and is distributed throughout the
body. Endosulfan is metabolized to lipophilic compounds and both the parent and
metabolites are found initially primarily in the kidney and liver and fatty tissue, with
distribution to other organs occurring over time. Endosulfan can induce
microsomal enzyme activity and is a nonspecific inducer of drug metabolism. In
sheep, approximately 1 percent of a single dose was recovered in milk. Females
may accumulate endosulfan more readily than males according to animal studies.
This may be causal in the higher toxicity seen in females (see Acute Toxicity
below) (ATSDR, 1993a).
5.3.5.3 Acute Toxicity—
Acute accidental or intentional ingestion of large amounts of endosulfan has
resulted in death in humans. However, available data are insufficient to estimate
a lethal dose of endosulfan in humans. Mice appear to be quite sensitive to
endosulfan's lethal effects with an LD50 of 7 mg/kg. In rats, exposed males and
females appear to have different sensitivities to the lethal effects of endosulfan
(e.g. oral LD50 values were 10-23 mg/kg in females and 40-125 mg/kg in males).
Insufficient data were available to determine whether differences in sensitivity to
lethal effects exist between males and females of species other than the rat.
Acute toxicity in humans and animals involve a large number of organ systems
(respiratory, cardiovascular, gastrointestinal, hematological, hepatic, renal). The
most prominent sign of acute overexposure to endosulfan in both humans and
animals is central nervous system stimulation (hyperactivity, tremors, decreased
respiration, convulsions) (ATSDR, 1993a)."
5.3.5.4 Chronic Toxicity—
IRIS provides an RfD of 6 x 10"3 mg/kg-d (IRIS 1999). The principal study on
which this RfD is based was a 2-yr feeding study in rats. Reduced body weight
gain in males and females, increased incidence of marked progressive
glomerulonephrosis, and blood vessel aneurysms in males were observed. The
LOAEL for systemic toxicity was 2.9 mg/kg-day in males and 3.8 mg/kg-d in
females. The NOAEL for systemic toxicity was 0.6 mg/kg-d in males and 0.7
mg/kg-d in females. The NOAEL of 0.6 mg/kg-d was divided by an uncertainty
factor of 100; 10 for intraspecies variability and 10 for interspecies extrapolation.
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5.3.5.5 Reproductive and Developmental Toxicity—
In a two-generation reproduction study in rats, no evidence of reproductive toxicity
was found at the highest dose tested of 6 mg/kg-d. The NOAEL for offspring
toxicity was 1 mg/kg-d based on increased pituitary and uterine weights at the
next higher dose of 6 mg/kg-d. A number of adverse effects were noted in a
developmental study in rats (increased incidence of misaligned sternebrae, extra
ribs, poor ossification). However, the study had a number of deficiencies and the
US EPA recommended that it be repeated. In a study in rabbits, no developmental
effects were noted at the highest dose tested of 1.8 mg/kg-d (IRIS, 1994).
5.3.5.6 Mutagenicity—
Results of mutagenicity assays of endosulfan are mixed, with multiple positive and
negative studies (ATSDR, 1993a; HSDB, 1993; IRIS, 1999). Endosulfan has
resulted in an increase in the percentage of aberrant colonies and the frequency
of gene convertants and revertants in yeast and was genetically effective without
activation. Longer duration of exposure increased effects (HSDB, 1993). In vivo
assays have found chromosomal aberrations and gene mutations in mice
(ATSDR, 1993a). However, some of these data may be suspect because some
formulations contained epichlorohydrin, a known genotoxic chemical, as a
stabilizer (ATSDR, 1993).
5.3.5.7 Carcinogenicity—
ATSDR has concluded that the available animal study data were negative or
inconclusive (ATSDR, 1993b). EPA has classified endosulfan in Group E
(evidence of noncarcinogenicity for humans) (U.S. EPA, 1999c).
5.3.5.8 Special Susceptibilities—
The limited toxicity data available for endosulfan suggest that several subgroups
of the population may be more susceptible to endosulfan exposure than the
general population. These subgroups include those with liver, kidney,
immunological, or blood diseases; compromised immune systems such as AIDS
patients, infants, and elderly people; hematologic disorders; seizure disorders;
and low protein diets (see below) (ATSDR, 1993a).
There is evidence from animal studies indicating that unborn and
neonates may be more susceptible to the toxic effects of
endosulfan because hepatic detoxification systems are immature
and therefore unable to metabolize xenobiotic substances
efficiently (ATSDR, 1993a).
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5.3.5.9 Interactive Effects-
Human anecdotal information suggests that endosulfan may act synergistically
with alcohol (ATSDR, 1993a). In rats, moderate protein deprivation doubled the
toxicity of endosulfan (ATSDR, 1993a).
Pentobarbital and endosulfan have demonstrated an interactive effect that is
probably related to microsomal enzyme activity. Endosulfan induces the mixed
function oxidase system (ATSDR, 1993a). Vitamin A inhibited the endosulfan-
induced activity of the mixed function oxidase system (ATSDR, 1993a).
5.3.5.10 Critical Data Gaps—
The increased susceptibility of female rats to endosulfan should be studied to
determine the underlying cause and to evaluate whether the effect occurs with
chronic species other than the rat.
Additional data are needed on the teratogenic and neurobehavioral effects during
development resulting from endosulfan exposure. Current data do not provide a
consistent picture nor do they explain underlying mechanisms of toxicity.
A joint team of scientists from ATSDR, NTP, and EPA have identified the following
data gaps: acute oral exposure studies, mechanisms of anemia-inducing effects,
reproductive system toxicity and related performance, developmental toxicity
studies, mechanisms of immunotoxicity, sensitive neurological function and
histological studies for long-term exposures, epidemiological studies,
pharmacokinetics of intermediate and chronic duration exposures, and studies
evaluating mechanisms underlying the differences in male and female toxicity.
No ongoing studies were identified for endosulfan (ATSDR, 1993a).
5.3.5.11 Summary of EPA Health Benchmarks-
Chronic Toxicity 6 x 10~3 mg/kg-d
Carcinogenicity Group E (no evidence of carcinogenicity).
5.3.5.12 Major Sources—
ATSDR (1993a), HSDB (1993), IRIS (1999), U.S. EPA (1993g).
5.3.6 Endrin
5.3.6.1 Background—
Endrin is an organochlorine pesticide whose registration was canceled in 1984
(U.S. EPA, 1993a).
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5.3.6.2 Pharmacokinetics—
Endrin, like the other organochlorine pesticides, is lipophilic. It bioaccumulates
and is distributed in fat, the liver, the brain, and kidneys and is rapidly metabolized
in mammals via oxidation of the methylene bridge. Metabolic products are
probably more toxic than endrin and the toxic entity has been hypothesized to be
12-ketoendrin. In humans, this compound is excreted directly in urine and feces
(ATSDR, 1990).
5.3.6.3 Acute Toxicity—
The primary target of endrin is the central nervous system (ATSDR, 1990).
5.3.6.4 Chronic Toxicity—
IRIS provides an RfD of 3 x 10'4 mg/kg-d based on a NOAEL of 0.025 mg/kg-d
from a 1969 chronic exposure dog study that identified mild histological effects in
the liver and occasional convulsions in study subjects exposed at the LOAEL of
0.05 mg/kg-d. Uncertainty factors of 10 each for inter- and intraspecies variability
were applied (IRIS, 1999).
OPP tox one-liners list a 1959 2-year dog feeding study with a LOAEL of 0.015
mg/kg-d based on hypersensitivity in the neck and shoulder area. Increased
erythropoiesis was noted at 0.125 mg/kg-d (U.S. EPA, 1993k). The LOAEL of
0.015 is within 1 order of magnitude of the LOAEL identified in the critical IRIS
study.
5.3.6.5 Reproductive and Developmental Toxicity—
No developmental effects were listed in the IRIS file for endrin (IRIS, 1999).
ATSDR listed a number of prenatal exposure studies that identified structural
abnormalities and neurotoxicity associated with endrin exposure. Structural
abnormalities have been observed in mice and hamsters exposed to endrin.
These include fused ribs and cleft palate at 5 mg/kg-d for 3 prenatal days and
webbed foot and open eye effects in hamster fetuses prenatally exposed for 1
day. Meningeocephaloceles in hamsters were caused by a single prenatal
exposure "above" 1.5 mg/kg and fused ribs "above" 5 mg/kg in hamsters. In
mice, a single prenatal exposure to 2.5 mg/kg caused an increase in open eyes.
Exencephaly and fused ribs were seen with one exposure at 9 mg/kg endrin. A
rat study reported no developmental effects with exposure to 0.45 mg/kg-d (it was
not clear if behavioral effects were evaluated) (ATSDR, 1990). The variation in
effects is probably due in part to the different prenatal periods during which
exposure occurred (see ATSDR, 1990). Reproductive outcome was adversely
affected in hamsters exposed to 1.5 mg/kg-d with decreased survival of pups (16
percent mortality) (ATSDR, 1990).
Nervous system effects are a significant concern with organochlorine exposure.
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In hamsters, abnormally increased pup activity in hamsters was observed with 1.5
mg/kg prenatal exposures for 9 days. The NOAEL for these behavioral effects
was 0.075 mg/kg-d (ATSDR, 1990). In rats, increased activity was seen with
prenatal exposure to 0.3 mg/kg-d (ATSDR, 1990). Abnormally increased activity
has been observed for other organochlorine pesticides (see DDT) and has been
associated with probable altered learning ability and permanent structural
changes to the brain.
As noted in the pharmacokinetics section above, endrin can accumulate in body
tissue; consequently, exposure occurring prior to pregnancy can contribute to the
overall maternal body burden and result in exposure to the developing individual.
As a result, it is necessary to reduce exposure to children and females with
childbearing potential to reduce overall body burden. If exposure is reduced
during pregnancy but has occurred prior to pregnancy, the pregnancy outcome
may be affected, depending on the timing and extent of prior exposure.
5.3.6.6 Mutagenicity—
In vitro assays of endrin suggest that it is not genotoxic. There were no in vivo
assay results located (ATSDR, 1990).
5.3.6.7 Carcinogenicity—
Insufficient information is available to determine the carcinogenic status of endrin.
EPA has classified endrin as a Group D carcinogen (not classifiable as to human
carcinogenicity). Some studies have yielded positive results and some studies that
reported negative results were considered to be inadequate (IRIS, 1999). Tumors
have been noted in the adrenal glands, pituitary glands, liver, mammary gland,
uterus, and thyroid in various studies and multiple species (IRIS, 1999). Endrin
is structurally related to a number of chemicals that are carcinogenic in test
animals, including chlordane, aldrin, dieldrin, heptachlor, and chlorendic acid
(IRIS, 1999). Because endrin has been classified as a Group D carcinogen, no
cancer potency has been listed by EPA.
5.3.6.8 Special Susceptibilities—
ATSDR has reported that children may be more sensitive to acute endrin
exposure than adults, based on effects observed in children during a poisoning
incident. Children appeared more susceptible to neurotoxic effects and have
exhibited convulsions. This is supported by results observed in experimental
animals where young rats were more susceptible than adults (ATSDR, 1990).
In addition, the skeletal and behavioral abnormalities associated with endrin
exposure in experimental animals indicate that prenatal exposure may generate
special risks.
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Based on animal studies, females may be more susceptible than males to endrin-
induced toxicity (ATSDR, 1990).
5.3.6.9 Interactive Effects-
Dietary pretreatment with endrin potentiates the hepatotoxicity of carbon
tetrachloride. MIXTOX has reported synergism between endrin and chlordane in
mice with gavage exposure (MIXTOX, 1992).
5.3.6.10 Critical Data Gaps—
A joint team of researchers from ATSDR, NTP, and EPA have identified the
following data gaps: human responses to acute, intermediate (14 to 365 days),
and chronic exposures; subchronic reproductive tests in various species;
immunotoxicity studies of animals and humans; human dosimetry studies;
pharmacokinetic studies; and studies of interspecies differences in metabolism
and toxicity (ATSDR, 1990).
5.3.6.11 Summary of EPA Health Benchmarks-
Chronic Toxicity 3 x 10~4 mg/kg-d
Carcinogenicity Group D (not classifiable).
5.3.6.12 Major Sources—
ATSDR (1990), IRIS (1999), U.S. EPA (1993k).
5.3.7 Heptachlor Epoxide
5.3.7.1 Background—
Heptachlor epoxide is a breakdown product of the organochlorine pesticides
heptachlor and chlordane and is a contaminant of both products. It is more toxic
than either parent compound (ATSDR, 1993b). Although most uses of heptachlor
were suspended in 1978 and chlordane was removed from the market in 1988
(U.S. EPA, 1993h), heptachlor epoxide continues to be a widespread contaminant
due to its relatively long half-life.
5.3.7.2 Pharmacokinetics—
Based upon animal and limited human data, heptachlor epoxide is absorbed
through the Gl tract and is found primarily in the liver, bone marrow, brain, and fat,
although it is distributed widely to other tissues as well. It is stored primarily in fat.
Fetal blood levels were approximately four times those measured in women.
Levels in human milk range from zero to 0.46 ppm (ATSDR, 1993b).
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Heptachlor epoxide has a very long half-life, particularly in adipose tissue. Human
tissue levels have correlated well to age, with 97 percent of North Texas residents
tested (ages 41 to 60) having measurable levels. Based on the Texas study,
heptachlor epoxide tissue levels have not decreased appreciably since the 1960s
(ATSDR, 1993b).
5.3.7.3 Acute Toxicity—
The LD50s for heptachlor range from 40 to 162 mg/kg in rodents (ATSDR, 1993b).
5.3.7.4 Chronic Toxicity—
IRIS provides an RfD of 1.3 x 10"5 mg/kg-d based on an LOAEL of 0.0125 mg/kg-
d from a 60-week dog feeding study reported in 1958. The critical effect was
increased liver-to-body-weight ratios in both males and females at the lowest dose
tested. Uncertainty factors of 10 each were applied for inter- and intraspecies
variability and the use of an LOAEL rather than a NOAEL (IRIS, 1999). No
additional uncertainty factors were applied for the use of a less-than-lifetime
study. The principal study is of low quality and there is low confidence in the RfD
(IRIS, 1999).
Animal studies have identified the following effects associated with heptachlor
(and subsequently heptachlor epoxide via metabolism) or heptachlor epoxide
directly: elevated bilirubin and white blood cell count, increased serum creatinine
phosphokinase levels suggestive of muscle damage, muscle spasms secondary
to CNS stimulation, adrenal gland pathology, and neurological disorders (ATSDR,
1993b). Significant changes in EEG patterns were found in female adult rats
exposed to 1 and 5 mg/kg-d for three generations (ATSDR, 1993b).
5.3.7.5 Reproductive and Developmental Toxicity—
A human study conducted in Hawaii was not considered adequate due to many
study design deficiencies (ATSDR, 1993b). In another epidemiological study of
women who had premature deliveries, significantly higher levels of heptachlor
epoxide and other organochlorine pesticides were detected in sera (ATSDR,
1993b).
A 1973 two-generation dog reproductive study identified a NOAEL of 0.025
mg/kg-d with an LOAEL of 0.075 mg/kg-d with liver lesions in pups. Other studies
with higher LELs based on a lethality endpoint are listed in the IRIS file. They
were not used in this evaluation due to insufficient information. The IRIS file notes
data gaps as rat and rabbit teratology studies (IRIS, 1999).
Exposure of adult rats to 6 mg/kg-d caused lens cataracts in 22 percent of the
adults, 6 to 8 percent of the F1 generation offspring, and 6 percent of the F2
generation offspring. A rat study with exposure to 0.25 mg/kg-d occurring 60 days
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prior to mating and during gestation resulted in severely reduced pup survival (15
percent) at 21 days postpartum (ATSDR, 1993b).
As noted in Section 5.3.7.2, heptachlor can accumulate in body tissue; con-
sequently, exposure occurring prior to pregnancy can contribute to the overall
maternal body burden and result in exposure to the developing individual. As a
result, it is necessary to reduce exposure to children and females with child-
bearing potential to reduce overall body burden. If exposure is reduced during
pregnancy but has occurred prior to pregnancy, the pregnancy outcome may be
affected, depending on the timing and extent of prior exposure.
A study of reproductive system toxicity with males and females dosed at 0.25
mg/kg-d prior to and during gestation found a significantly decreased pregnancy
rate among exposed animals. Based on specific fertility tests, it was determined
that males were most likely affected and that sperm were probably killed (ATSDR,
1993b). Another reproductive system toxicity study with doses at and above
0.075 mg/kg-d resulted in the failure of animals to reproduce. There were serious
deficiencies in this study (ATSDR, 1993b).
5.3.7.6 Mutagenicity—
Mixed results have been obtained in mutagenicity assays of heptachlor epoxide.
5.3.7.7 Carcinogenicity—
Heptachlor epoxide is classified as a probable human carcinogen (B2) by EPA
based on oral studies in animals. The oral cancer slope factor is 9.1 per mg/kg-d.
This value is based on the geometric mean of several studies that identified liver
carcinomas (IRIS, 1999). Five structurally related compounds have produced
tumors in mice and rats: chlordane, aldrin, dieldrin, heptachlor, and chlorendic
acid (IRIS, 1999).
Statistically significant increases in adenomas and carcinomas of the thyroid were
found in female rats. Some researchers discounted the results due to the low
incidence and known variability in the control population (ATSDR, 1993b).
Heptachlor (and consequently heptachlor epoxide) exposures have been asso-
ciated with cerebral gliosarcoma in children exposed prenatally. Multiple chromo-
somal abnormalities were also identified in the tumor cells. It was not determined
whether the effects were caused by environmental or familial factors (ATSDR,
1993b).
5.3.7.8 Special Susceptibilities-
Based on the toxicity data reviewed above, individuals with diseases or disorders
of the following systems may be at greater risk than the general population: liver,
hematopoietic, musculoskeletal, neurological, and adrenal gland. ATSDR has
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noted that preadolescent children may be more susceptible due to their greater
rate of glutathionine turnover (ATSDR, 1993b). In addition, children exposed
prenatally may be at higher risk, based on the results of developmental toxicity
studies.
5.3.7.9 Interactive Effects—
Heptachlor induces the mixed function oxidase system. No specific interactive
effects have been noted.
5.3.7.10 Critical Data Gaps—
The IRIS file notes data gaps as rat and rabbit teratology studies (IRIS, 1999).
A joint team of scientists from EPA, NTP, and ATSDR have identified the following
data gaps: a model to describe the relationship between tissue and blood levels
and exposure in humans, chronic oral exposure effects in humans,
epidemiological and in vivo animal genotoxicity studies, developmental and
reproductive toxicity studies and neurotoxicity and immunotoxicity studies in
animals, and pharmacokinetic studies (ATSDR, 1993b).
5.3.7.11 Summary of EPA Health Benchmarks-
Chronic Toxicity 1.3 x 10~5 mg/kg-d
Carcinogenicity 9.1 per mg/kg-d.
5.3.7.12 Major Sources—
ATSDR (1993b), IRIS (1999).
5.3.8 Hexachlorobenzene
5.3.8.1 Background—
Hexachlorobenzene was used as a fungicide on seeds of onions, sorghum,
wheat, and other grains until 1984. It was also used in pyrotechnics and as a
chemical intermediate but is no longer used commercially in the United States
(ATSDR, 1996b).
5.3.8.2 Pharmacokinetics—
Hexachlorobenzene is persistent in the body, accumulating preferentially in fat
and tissues with a high lipid content, because of its lipophilic nature. It is found
in human breast milk (ATSDR, 1996b), which may be a significant route of
exposure for young children. Hexachlorobenzene is also readily transferred
through the placenta from the mother to the fetus in animal experiments.
Hexachlorobenzene is very slowly converted by microsomal enzymes in the liver
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to its major metabolites, pentachlorophenol, pentachlorothiphenol, and
pentachlorobenzene, which are mainly excreted in the urine.
5.3.8.3 Acute Exposure-
Acute exposure studies in animals indicate a relatively low acute toxicity with
LD50s between 1,700 and 4,000 mg/kg (ATSDR, 1996b). Exposure to
hexachlorobenzene does not appear to cause the acute neurological effects
observed with the organochlorines that have been used as insecticides (e.g.,
DDT). Based on animal studies, the following systems are adversely affected
following acute exposure: liver, kidney, hematological, endocrine, and dermal
(ATSDR, 1996b).
5.3.8.4 Chronic Toxicity—
Hexachlorobenzene exposure of a large number of people in Turkey occurred
between 1955 and 1959 due to consumption of contaminated grain. No precise
exposure estimates are available for children or adults in this episode; it is likely
that exposures occurred over a continuum, with some individuals consuming
much higher levels than others. Researchers have estimated relatively low
exposure levels occurred over several years as a result of consumption (50 to 200
mg/d). These exposure levels are approximately 0.7 to 2.9 mg/kg-d for a 70-kg
individual. It should be emphasized that the exposure estimates are unverified
(ATSDR, 1996b).
The following effects have been associated with hexachlorobenzene exposure in
individuals exposed chronically via contaminated bread (Turkey): shortening of the
digits due to osteoporosis, painless arthritis, decreased uroporphyrin synthase
levels, muscle weakness, rigidity and sensory shading, thyroid enlargement, and
histopathological changes in the liver often accompanied by skin lesions (ATSDR,
1996b). These effects were also observed in numerous animal studies (See
discussion under Section 5.3.8.5 also.)
The hepatic system appears to be the most sensitive systemic endpoint for
hexachlorobenzene exposure, IRIS provides an RfD value of 8 x 10~4 mg/kg-d
based on a NOAEL of 0.08 mg/kg-d in a lifetime rat study. An uncertainty factor
of 100 was applied; 10 for interspecies and 10 for intraspecies variability.
Numerous other studies identified NOAELs in the same numerical range, so the
confidence in the database is rated as high. The IRIS file notes that the sensitive
endpoint of porphyria, which is an effect noted in exposed human populations,
was not evaluated in the critical animal study, so the confidence in the RfD is
rated as medium (IRIS, 1999).
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5.3.8.5 Reproductive and Developmental Toxicity—
Lactational exposure to hexachlorobenzene is of significant concern, based on the
rapid transfer of the chemical through breast milk and effects observed in children
of exposed mothers in a contamination incident in Turkey. In a study of nursing
infants, blood levels of hexachlorobenzene were two to five times that of their
mothers; tissue levels were higher as well. A study of monkeys found that the
concentration in milk was 17 times higher than that in maternal serum (ATSDR,
1996b). Young children (under 1 year) of lactating mothers who were exposed
via contaminated bread had an extremely high mortality rate. Skin lesions,
weakness, and convulsions were reported in these infants. Although adults were
also adversely affected, children appeared to be at higher risk. The maternal
exposure was roughly estimated to be 0.7 to 2.9 mg/kg-d (ATSDR, 1996b).
Among slightly older children (average age of 7), exposure via food resulted in the
development of small or atrophied hands and fingers, short stature, pinched
faces, osteoporosis in the hands, and other arthritic changes. Exposure was
estimated to be approximately 0.7 to 2.9 mg/kg-d (ATSDR, 1996b).
It is known that hexachlorobenzene can cross the human placenta; however, no
data were available on effects resulting from prenatal exposure in humans. Very
limited information is available on experimental animals. Cleft palate and kidney
abnormalities were observed in one study in a single litter and fetus at 100 mg/kg-
d (ATSDR, 1996b). In another study, the survivability of prenatally exposed rats
was significantly reduced at 2 mg/kg-d (estimated from ppm with conversion factor
of 0.05 mg/kg per 1 ppm diet for rats). Death was attributed to maternal body
burden and cumulative lactational exposure (ATSDR, 1996b). Alterations in
immune function levels were reported in pre- and postnatally exposed rats at 4
mg/kg (ATSDR, 1996b).
As noted above, hexachlorobenzene accumulates in body tissue; consequently,
exposure occurring prior to pregnancy can contribute to the overall maternal body
burden and result in exposure to the developing individual. As a result, it is
necessary to reduce exposure to children and women with childbearing potential
to reduce overall body burden. If a female has been exposed to hexachloro-
benzene, even if exposure is reduced during pregnancy, the outcome of that
pregnancy may be affected, depending on the timing and extent of prior exposure.
5.3.8.6 Mutagenicity—
The results of mutagenicity studies on hexachlorobenzene are mixed (IRIS,
1999). Hexachlorobenzene was negative in dominant lethal studies (in vivo) at
doses from 60 to 221 mg/kg (ATSDR, 1996b).
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5.3.8.7 Carcinogenicity—
Carcinogenic assays of hexachlorobenzene in animals have identified an
increased incidence of multiple tumor types including hepatomas, hemangioendo-
theliomas, liver, and thyroid tumors in multiple species. EPA developed a cancer
potency of 1.6 mg/kg-d based on liver carcinoma in female rats exposed via diet.
In support of this value, cancer potencies were calculated for 14 different data
sets; the results were within 1 order of magnitude. Hexachlorobenzene is
classified as a probable human carcinogen (B2) based on the results of animal
studies (IRIS, 1999).
Follow-up studies of exposure victims in Turkey have not identified cancers in the
25- and 20- to 30-year exposure cohorts; however, ATSDR suggests that the
enlarged thyroids noted in members of these groups have not been sufficiently
investigated (ATSDR, 1996b). It should also be noted that most cancers have
multiple-decade latency periods and often occur in the later part of life.
Consequently, it will not be possible to assess the carcinogenic impact of
exposures in Turkey for some time.
5.3.8.8 Special Susceptibilities—
ATSDR has concluded that young children are susceptible to hexachlorobenzene
exposure based on human poisoning episodes. Exposure led to permanent
debilitating effects. Both human and animal data suggest that the risk of exposure
to nursing infants may be greater than the risk to their mothers (ATSDR, 1996b).
Based on the toxicity data reviewed above, individuals with liver disease may be
at greater risk than the general population.
5.3.8.9 Interactive Effects—
Hexachlorobenzene induces microsomal enzymes. Pentachlorophenol increases
the porphyrinogenic effects of hexachlorobenzene. Hexachlorobenzene
potentiated the thymic atrophy and body weight loss caused by 2,3,7,8-TCDD.
A 50 percent food deprivation increased liver hypertrophy and microsomal
enzyme induction by hexachlorobenzene (ATSDR, 1996b).
5.3.8.10 Critical Data Gaps—
A joint team of scientists from EPA, NTP, and ATSDR have identified the study
following data gaps: human carcinogenicity, in vivo and in vitro genotoxicity,
animal reproductive toxicity, animal developmental toxicity, immunotoxicity studies
in humans, and pharmacokinetics (ATSDR, 1996b). Information is needed to
develop a model that can be used to estimate the relationship between maternal
intake, human milk concentration, and adverse effects in infants.
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5.3.8.11 Summary of EPA Health Benchmarks-
Chronic Toxicity 8 x 10~4 mg/kg-d
Carcinogenicity 1.6 per mg/kg-d.
5.3.8.12 Major Sources—
ATSDR (1996b), IRIS (1999).
5.3.9 Lindane (y-hexachlorocyclohexane)
5.3.9.1 Background—
Lindane is an organochlorine pesticide that is comprised of isomers of
hexachlorocyclohexane, with the y isomer constituting the major (>99 percent)
component. There appears to be some difference in toxicity of the various
hexachlorocyclohexane isomers (U.S. EPA, 1993a). The following data assume
that lindane can be defined as the y isomer. Lindane is used primarily for
controlling wood-inhabiting beetles and as a seed treatment. Lindane is also used
as a prescription pharmaceutical to control head lice and mites (scabies) in
humans.
5.3.9.2 Pharmacokinetics—
Lindane is readily absorbed by the Gl tract following oral exposure. Distribution
is primarily to the adipose tissue but also to the brain, kidney, muscle, spleen,
adrenal glands, heart, lungs, blood, and other organs. It is excreted primarily
through urine as chlorophenols. The epoxide metabolite may be responsible for
carcinogenic and mutagenic effects (ATSDR, 1994c).
Male exposure to lindane through the environment results in accumulation in
testes and semen in addition to the tissues listed above (ATSDR, 1994c). See
also a discussion in Section 5.3.9.5 of the accumulation of lindane by pregnant
women.
5.3.9.3 Acute Toxicity—
The estimated human lethal dose is 125 mg/kg (HSDB, 1993). Occupational and
accidental exposures in humans have resulted in headaches, vertigo, abnormal
EEG patterns, seizures, and convulsions. Death has occurred primarily in
children.
5.3.9.4 Chronic Toxicity—
IRIS provides an RfD of 3 x 10'4 mg/kg-d based on a NOAEL of 0.33 mg/kg-d
from a subchronic rat study that found liver and kidney toxicity at higher doses.
Uncertainty factors of 10 each for inter- and intraspecies variability and the use
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of a less-than-lifetime study were applied (IRIS, 1999). The confidence in the
principal study, database, and RfD are rated as medium. A recently completed
2-year study is under evaluation and may provide additional information regarding
toxicity (U.S. EPA, 1993i). Liver damage has been observed in many animal
studies and appears to be the most sensitive effect (U.S. EPA, 1993i). Immune
system effects have been observed in humans exposed via inhalation and in
orally dosed animals. A 5-week study in rabbits found immunosuppression at 1
mg/kg-d (ATSDR, 1994c).
Most observed effects in humans exposed accidentally to lindane are
neurological. Behavioral effects have also been noted in many studies on
experimental animals, and at relatively high levels seizures were reported. More
subtle behavioral effects were noted at an LOAEL of 2.5 mg/kg-d with 40 days of
exposure in rats. No NOAELwas reported (ATSDR, 1994c).
5.3.9.5 Reproductive and Developmental Toxicity—
Two developmental toxicity studies in rats and rabbits both identified a NOAEL of
10 mg/kg (no effects were described for higher doses). A three-generation rat
study found no adverse reproductive effects at 5 mg/kg-d, the highest dose tested
(U.S. EPA, 1993i). A recent mouse study found increased resorptions at 5 mg/
kg-d. Studies in rats and mice have found increased incidence of extra ribs at 5
to 20 mg/kg-d (ATSDR, 1994c). There are multiple studies showing pre- and
postimplantation fetotoxicity and skeletal abnormalities resulting from prenatal
exposure at higher doses (HSDB, 1993).
Lindane accumulates in the fatty tissue of pregnant (and nonpregnant) women
where it can be transferred to the fetus through the placenta and to infants
through breast milk. Human milk concentrations are approximately five to seven
times greater than maternal blood levels. Concentrations in maternal blood are
proportional to the length of time overwhich exposure occurred, with olderwomen
having higher blood levels. During pregnancy, the lindane concentration in blood
from fetal tissue, uterine muscle, placenta, and amniotic fluid was higher than
levels in maternal adipose tissue, and blood serum levels increased during
delivery (ATSDR, 1994c). There is little information on the effects of exposure
during lactation. One study (dose unspecified) in rats indicated that exposure
during gestation and lactation did not cause developmental effects; however, this
is not consistent with other studies that found effects associated with gestational
exposure.
Based on what is known regarding the transfer of lindane into human milk, nursing
infants must be considered at some risk if their mothers have been exposed to
significant amounts of lindane (lindane is a lipid-seeking chemical). Additional
information is needed to characterize the relationship between maternal intake,
body burden (blood or adipose levels), milk concentrations, and adverse effects.
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Multiple studies have reported that lindane exposure (as measured by body tissue
level of lindane) is associated with premature labor and spontaneous abortions.
The causal relationship has not been established for this action (ATSDR, 1994c);
however, the reproductive system effects discussed in Section 5.3.9.4 (bio-
chemical changes in uterine, cervical, and vaginal tissues and antiestrogenic
effects) may be involved.
As noted above, lindane accumulates in body tissue; consequently, exposure
occurring prior to pregnancy can contribute to the overall maternal body burden
and result in exposure to the developing individual. As a result, it is necessary to
reduce exposure to children and women with childbearing potential to reduce
overall body burden. If exposure is reduced during pregnancy but has occurred
prior to pregnancy, the pregnancy outcome may be affected, depending on the
timing and extent of prior exposure.
Two recent reproductive studies in rats found adverse effects on the male
reproductive system. In a 7-wk study, decreased sperm counts were noted at 50
mg/kg-d and, in a 180-d study, seminiferous tubular degeneration was noted at
6 mg/kg-d with a NOAEL of 3 mg/kg-d. An older study had identified the same
effects at 64.6 mg/kg-d in a 3-mo study. Experimental data indicate that the
female reproductive system may also be altered by lindane exposure. A study of
rats found uterine, cervical, and vaginal biochemical changes at 20 mg/kg-d in a
30-d study. Antiestrogenic effects were found at 20 mg/kg-d in female rats in a
15-wk study with a NOAEL of 5 mg/kg-d. This action was also found in two other
recent studies (ATSDR, 1994c).
5.3.9.6 Mutagenicity—
In animals, ingestion of technical-grade hexachlorocyclohexane-induced dominant
lethal mutations in mice. Studies found that lindane binds to mouse liver DMA at
a low rate. Based on a review of genotoxicity studies, ATSDR concluded that
lindane "has some genotoxic potential, but the evidence for this is not conclusive"
(ATSDR, 1994c).
5.3.9.7 Carcinogenicity—
Lindane has been classified as Group B2/C (probable/possible human
carcinogen) (U.S. EPA, 1999c) and a cancer potency of 1.3 per mg/kg-d has been
listed (HEAST, 1997). Lindane's related isomers, alpha and beta
hexachlorocyclohexane, are classified as probable human carcinogens and have
cancer potencies similar to that of lindane. In addition to tumors identified in
experimental animals, human study data indicate that this chemical may cause
aplastic anemia (U.S. EPA, 1993a).
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5.3.9.8 Special Susceptibilities—
ATSDR has recommended that pregnant and/or lactating women should not be
exposed to lindane. The potential for premature labor and spontaneous abortion
is noted (ATSDR, 1994c). People with epilepsy, cerebrovascular accidents, or
head injuries who have lower thresholds for convulsions may be at greater risk of
lindane-induced CMS toxicity and seizures. Also, individuals with protein-deficient
diets, liver or kidney disease, or immunodeficiencies may be at greater risk from
lindane exposure than the general population (ATSDR, 1994c).
Children may also be at greater risk from lindane exposure because of the
immaturity of their immune and nervous systems. ATSDR has cautioned that:
Infants and children are especially susceptible to immuno-
suppression because their immune systems do not reach maturity
until 10 to 12 years of age (ATSDR, 1994c).
5.3.9.9 Interactive Effects-
High- and low-protein diets and vitamin A and C deficiencies increased the toxicity
of lindane in experimental animals. Vitamin A supplements decreased toxicity.
Cadmium inhibited the metabolism of lindane. Combined cadmium and lindane
exposure caused significant embryotoxic and teratogenic effects in rats at
dosages that caused no effects when administered alone. Exposure to the a, P,
and 5 hexachlorocyclohexane isomers may reduce the neurotoxic effects of
lindane (ATSDR, 1994c).
MIXTOX has reported mixed results for studies of lindane and chlordane, lindane
and hexachlorobenzene, lindane and toxaphene, and lindane and mirex
interactions, including inhibition, no effect, and potentiation forthese combinations
in rodents exposed via gavage (MIXTOX, 1992).
5.3.9.10 Critical Data Gaps—
As discussed above, effects on both the male and female reproductive systems
have been evaluated in short-term studies. Evaluation of these effects in a
longer-term study and identification of the underlying mechanisms of toxicity
would provide information needed for a more complete evaluation of toxicity and
dose-response dynamics. Additional information is also needed, as noted in
Section 5.3.9.5, on the potential for exposure via lactation and on mechanisms
and dose-response for premature labor and spontaneous abortion.
ATSDR has identified data gaps that include chronic duration oral studies; in vivo
genotoxicity tests; reproductive, developmental immunotoxicity, and neurotoxicity
studies; human studies correlating exposure levels with body burdens of lindane
and with specific effects; and pharmacokinetic studies (ATSDR, 1994c).
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5.3.9.11 Summary of EPA Health Benchmarks-
Chronic Toxicity 3 x 10~4 mg/kg-d
Carcinogenicity 1.3 per mg/kg-d.
5.3.9.12 Major Sources—
ATSDR (1994c), HSDB (1993), IRIS (1999).
5.3.10 Mirex
5.3.10.1 Background—
Mirex was used as both an organochlorine pesticide and fire retardant from the
late 1950s until 1975 (U.S. EPA, 1993a). A major use of mirex was for the control
of ants, particularly fire ants in the southern United States. Mirex has the potential
to concentrate many thousandfold in food chains (Hayes and Laws, 1991).
5.3.10.2 Pharmacokinetics—
Mirex is a lipophilic compound and is readily taken up in fat tissue. The highest
residues were found in fat and the liver. Based on a study in cows, it is also found
in milk. At 0.01- and 1-ppm dietary exposure for 32 weeks, cows' milk levels were
0.01 to 0.08 ppm (U.S. EPA, 1993m).
No clear data on half-life in humans were found; however, studies in primates
found that 90 percent of the original dose was retained in fat after 106 days. The
researchers predicted that mirex had an extremely long half-life in monkeys.
Based on this, mirex would be expected to have a very long half-life in humans.
5.3.10.3 Acute Toxicity—
Acute hepatic effects have been observed in experimental animals. These may
result from the following cytological effects: disaggregated ribosomes, glycogen
depletion, formation of liposomes, and proliferation of smooth endoplasmic
reticulum (U.S. EPA, 1993m).
5.3.10.4 Chronic Toxicity—
IRIS lists a chronic exposure RfD of 2 x 10~4 mg/kg-d for mirex based on a NOAEL
of 0.07 mg/kg-d from a chronic (2-year) dietary rat study. Effects noted in the
study at higher doses were: splenic fibrosis, nephropathy, renal medullary
hyperplasia, multiple types of liver damage, and cystic follicles of the thyroid. The
RfD is based on the latter two critical effects. Uncertainty factors of 10 each were
applied for inter- and intraspecies variability and a factor of 3 was applied for lack
of a complete database (multigenerational data on reproductive effects and
cardiovascular toxicity data). The IRIS file also indicates that effects on the testes
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(testicular degeneration, hypocellularity, and depressed spermatogenesis), which
were noted in other studies, may not have been detected in the critical study
because of age-related degenerative changes in the study animals (IRIS, 1999).
5.3.10.5 Reproductive and Developmental Toxicity—
Studies in animals suggest that both male and female reproductive systems are
adversely affected by mirex. Acute exposure of male rats to 6 mg/kg-d mirex daily
for 10 days decreased their fertility significantly. Although residues of mirex were
found in the testes of the 6-mg/kg-d dose-group males, this did not affect
reproduction parameters in subsequent mating trials. The authors attributed the
observed decrease in the incidence of pregnancy in females mated with males in
this dose group to a subclinical toxic effect as suggested by reduction in body
weight gain in the dosed males (ATSDR, 1995a).
In a 28-day dietary study, decreased sperm count was noted in male rats at
dosages as low as 0.025 mg/kg-d; testicular degeneration was observed at
dosage levels of 2.5 and 3.7 mg/kg-d. However, mirex fed to rats at 1.3 to 3.1
mg/kg-d for two generations resulted in no decrease in fertility. In contrast,
females given 1.8 to 2.8 mg/kg-d for two generations produced a decreased
number of litters. Administration of 0.25 mg/kg-d to male and female rats for 91
days prior to mating and then through lactation resulted in decreased mating and
litter size (ATSDR, 1995a).
Exposure of maternal rats and mice during gestation resulted in increases in
resorptions and stillbirths and decreases in postnatal viability at doses as low as
1.25 mg/kg-d when administered from gestation days 4 through 22. Examination
of fetuses at the end of gestation showed increases in the incidence of edematous
fetuses and fetuses with cardiac arrhythmia; the incidence was slightly increased
at doses as low as 0.1 mg/kg-d. Additional effects were reported in a few studies
and included enlarged cerebral ventricles; undescended testes; cleft palate; short
tail; decreased skeletal ossification, fetal weight, and liver and kidney weights; and
liver and thyroid lesions. Cataracts were also observed in offspring in several
studies from pre- and postnatal exposures (ATSDR, 1995a).
5.3.10.6 Mutagenicity—
Most genotoxicity tests reported in the tox one-liners are bacterial assays and are
negative (U.S. EPA, 1993m). A dominant lethal mutagenicity test in rats (in vivo)
found a decreased incidence of pregnancy at 6 mg/kg-d with a NOEL of 3 mg/kg-
d. Exposure took place over 10 days prior to mating. However, parameters
indicative of dominant lethality were unaffected by treatment (ATSDR, 1995a)
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5.3.10.7 Carcinogenicity—
A marked increased incidence in neoplastic nodules in the liver of both male and
female rats was observed in a 2-year feeding study with mirex(NTP, 1990). This
effect was noted at doses of 0.7 mg/kg-d and above in males and at 3.8 mg/kg-d
and above in females. In addition, increased tumors of the adrenal gland in male
rats and mononuclear cell leukemias in female rats were observed. EPA's Office
of Pesticide Programs has classified mirex as Group B2 (probable human
carcinogen) (HEAST, 1997). In addition, NTP considers mirex as "reasonably
anticipated to be a human carcinogen" based on sufficient evidence of
carcinogenicity in experimental animals (NTP, 2000).
5.3.10.8 Special Susceptibilities-
Juveniles may be more susceptible than adults based on the results of animal
studies. At 60 ppm (approximately 3 mg/kg-d), adult mice exposed for 15 days
experienced only weight loss; this level was lethal for young mice (Hayes and
Laws, 1991).
Based on a review of the toxicity data above, individuals with diseases or
disorders of the following organ systems may be at higher risk than the general
population: kidney, liver, spleen, thyroid, parathyroid, cardiovascular, and male
reproductive. Due to the developmental toxicity observed in experimental
animals, prenatal exposure and lactation exposure may pose a risk to children.
The possibility exists that newborn children may also develop cataracts if exposed
to mirex shortly after birth (ATSDR, 1995a).
5.3.10.9 Interactive Effects—
Mirex induces the mixed function oxidase system. No specific interactive effects
have been noted.
MIXTOX reports mixed results for interactions between lindane and mirex and for
Aroclor 1254 and mirex. Other studies of Aroclor and mirex have not found
interactive results (MIXTOX, 1992).
5.3.10.10 Critical Data Gaps-
Additional information is needed on the developmental effects of mirex to identify
a NOAEL for sensitive developmental toxicity endpoints so that a well-founded
exposure limit for developmental effects can be determined. In a related area, the
mutagenicity data indicate a potential mutagenic effect based on in vivo studies.
A better understanding of the relationship between the results of these types of
studies and mutagenic effects in the human population is needed. The chronic
exposure toxicity studies do not provide consistent results. Additional clarification
of the NOAELs for sensitive endpoints in this area is needed.
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5.3.10.11 Summary of EPA Health Benchmarks-
Chronic Toxicity 2 x 10~4 mg/kg-d
Carcinogenicity Group B2 (probable human carcinogen).
5.3.10.12 Major Sources—
ATSDR (1995a), Hayes and Laws (1991), IRIS (1999), U.S. EPA (1993m).
5.3.11 Toxaphene
5.3.11.1 Background—
Toxaphene is an organochlorine pesticide that is comprised of a mixture of at
least 670 chlorinated camphenes. Toxaphene was probably the most heavily
used pesticide in the United States during the 1970s after DDT was banned. It
was banned for most uses in 1982; all uses were banned in 1990. However, due
to its relatively long half-life, it persists in the environment. The soil half-life is
approximately 1 to 14 years (HSDB, 1993).
5.3.11.2 Pharmacokinetics—
The components of toxaphene are metabolized in mammals via dechlorination,
dehydrodechlorination, and oxidation, primarily through the action of the mixed
function oxidase system and other hepatic microsomal enzymes. Conjugation may
occur but is not a major route of metabolism. Each component of toxaphene has
its own rate of biotransformation, making the characterization of toxaphene
pharmacokinetics complex. Some components of toxaphene are highly lipophilic
and poorly metabolized; these components may accumulate in body fat (ATSDR,
1996c).
5.3.11.3 Acute Toxicity—
Acute high-level exposures to toxaphene and toxaphene-contaminated food have
resulted in death in adults and children with an estimated minimum lethal dose of
2 to 7 g, which is equivalent to 29 to 100 mg/kg for an adult male. LD50 values in
rats were 80 mg/kg for females and 90 mg/kg for males. Transient liver and
kidney effects, and periods of memory loss have been observed in humans after
single large oral exposures. In animals, the most sensitive organ is the liver.
Toxicity to the central nervous system, kidney, and adrenal glands have also been
observed (ATSDR, 1996c).
5.3.11.4 Chronic Toxicity—
IRIS does not provide a discussion of chronic effects of exposure to toxaphene
or an RfD (IRIS, 1999). An RfD of 2.5 x 10"4 mg/kg-d is listed in the Office of
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Pesticide Program's Reference Dose Tracking Report (U.S. EPA, 1997c) and has
been agreed upon by the Office of Pesticide Programs and the Office of Water.
Chronic exposure to toxaphene may result in damage to the following organ
systems: liver, kidney, adrenal, immunological, and neurological. Chronic
exposure to toxaphene may cause hormonal alterations. A study on chronic
exposures found increased levels of hepatic metabolism of the hormones estradiol
and estrone and a decrease in their uterotropic action. Some adverse effects of
toxaphene that do not occur with a single exposure may result from repeated
exposures. Exposures at 0.06 mg/kg-d over 5 weeks caused adrenal hormone
reductions, whereas a single dose of 16 mg/kg did not cause effects.
5.3.11.5 Reproductive and Developmental Toxicity—
Women exposed to toxaphene by entering a field that had recently been sprayed
with the chemical exhibited a higher incidence of chromosomal aberrations in
cultured lymphocytes than did unexposed women. Dermal and inhalation were
the probable routes of exposure; however, the exposure was not quantified
(ATSDR, 1996c). Animal study results suggest that toxaphene does not interfere
with fertility in experimental animals at the doses tested (up to 25 mg/kg-d)
(ATSDR, 1996c).
Adverse developmental effects, including immunosuppressive and behavioral
effects, were noted in experimental animals at levels below those required to
induce maternal toxicity. Immunosuppression (reduction in macrophage levels,
cell-mediated immunity, and humoral immunity) was observed in test animals
exposed during gestation and nursing as were alterations in kidney and liver
enzymes and delayed bone development. Other adverse effects noted in
offspring of maternally exposed individuals included histological changes in the
liver, thyroid, and kidney (ATSDR, 1996c).
Toxaphene is known to be rapidly conveyed into breast milk after maternal
exposure to the chemical. The half-life of toxaphene in milk has been estimated
at 9 days.
As noted above, toxaphene accumulates in body tissue; consequently, exposure
occurring prior to pregnancy can contribute to the overall maternal body burden
and result in exposure to the developing individual. Therefore, it is necessary to
reduce exposure to children and women with childbearing potential to reduce
overall body burden.
Depending on the timing and extent of an individual's prior exposure to
toxaphene, the outcome of pregnancy may be affected even if exposure during
pregnancy is reduced.
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5.3.11.6 Mutagenicity—
Changes in human genetic material have been noted in workers exposed to
toxaphene (HSDB, 1993). There are also numerous positive mutagenicity assays
of toxaphene: the Ames test, sister chromatid exchange, chromosomal
aberrations in toxaphene-exposed humans, and forward mutation assays. The
implications of this for human germ cells are not known. One assay designed to
assess the effects of dominant lethal effects on implantations in mice yielded
negative results. Some data suggest that the polar fraction of toxaphene may be
more mutagenicthan the nonpolar fraction (ATSDR, 1996c; HSDB, 1993).
5.3.11.7 Carcinogenicity—
Toxaphene is classified as a probable human carcinogen (B2) by EPA based on
oral studies in animals (IRIS, 1999). No conclusive human epidemiological studies
are available for toxaphene (ATSDR, 1996c). Oral administration of toxaphene
resulted in an increased incidence of hepatocellular carcinomas and neoplastic
nodules in mice, and thyroid tumors in rats (IRIS, 1999). The cancer potency is
1.1 per mg/kg-d, based on liver tumors in experimental animals (IRIS, 1999).
Toxaphene has recently been observed to have estrogenic effects on human
breast cancer estrogen-sensitive cells (Soto et al., 1994). Xenoestrogens have
been hypothesized to have a role in human breast cancer (Davis etal., 1993). In
addition to potential carcinogenic effects, toxaphene may also cause disruption
of the endocrine system due to its estrogenic activity (Soto et al., 1994).
5.3.11.8 Special Susceptibilities—
A protein-deficient diet may increase the toxicity of toxaphene approximately
threefold based on an LD50 study in rats (ATSDR, 1996c). Individuals with latent
or clinical neurological diseases, such as epilepsy or behavioral disorders, may
be at higher risk for toxaphene toxicity. In addition, children may be especially
susceptible to toxaphene-induced neurotoxicity based on early reports of acute
ingestion toxicity (ATSDR, 1996c).
Other individuals who may be at higher risk are those with diseases of the renal,
nervous, cardiac, adrenal, and respiratory systems. Individuals using certain
medications are also at potential risk due to the induction of hepatic microsomal
enzymes by toxaphene (discussed further in the following section).
5.3.11.9 Interactive Effects-
Metabolism of some drugs and alcohol may be affected by toxaphene's induction
of hepatic microsomal enzymes. This was observed in a man using warfarin as
an anticoagulant while he used toxaphene as an insecticide. The effectiveness
of the drug was reduced because toxaphene's induction of microsomal enzymes
increased the drug's metabolism (ATSDR, 1996c).
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Based on acute studies in animals and anecdotal reports of acute exposure in
humans, exposure to chemicals that increase microsomal mixed-function oxidase
systems (e.g., lindane) are likely to reduce the acute toxicity of other chemicals
detoxified by the same system (e.g., toxaphene) because the system is
functioning at a higher than normal level. Toxaphene, in turn, may reduce the
acute toxicity of chemicals that require this system for detoxification (ATSDR,
1996c).
5.3.11.10 Critical Data Gaps—
The following data gaps have been identified for toxaphene: mammalian germ cell
genotoxicity, studies that investigate sensitive developmental toxicity endpoints
including behavioral effects, epidemiological and animal studies of immunotoxicity,
long-term neurotoxicity studies in animals using sensitive functional and
neuropathological tests and behavioral effects on prenatally exposed animals,
epidemiological studies evaluating multiple organ systems, and pharmacokinetic
studies (ATSDR, 1996c).
5.3.11.11 Summary of EPA Health Benchmarks-
Chronic Toxicity 2.5 x 10~4 mg/kg-d
Carcinogenicity 1.1 per mg/kg-d.
5.3.11.12 Major Sources—
ATSDR (1996c), HSDB (1993), IRIS (1999).
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5.4 ORGANOPHOSPHATE PESTICIDES
Please note that these analytes are currently undergoing reassessment by the
EPA under the provisions of the Food Quality Protection Act of 1996. This
reassessment may result in changes in the RfD values. Contact EPA for the most
current information.
5.4.1 Chlorpyrifos
5.4.1.1 Background—
Chlorpyrifos is an organophosphate insecticide first registered in 1965 and used
throughout the United States. Chlorpyrifos is used to control foliar and soil insects
for a wide variety of crops. While most use is agricultural, significant amounts of
Chlorpyrifos are used in urban settings for termite control and commercial
landscape maintenance and pest control. Chlorpyrifos formulations (e.g.,
Dursban) are also used by the general public for home, lawn, and garden insect
control.
5.4.1.2 Pharmacokinetics—
Chlorpyrifos accumulates in fat and has a longer half-life in fatty tissues than in
other tissues. It has been detected in cows' milk (HSDB, 1993) and would be
expected to occur in human milk of exposed mothers. This is of concern because
organophosphates may have a higher toxicity for immature individuals than adults
(e.g., malathion was more toxic to juveniles in three species tested) (U.S. EPA,
1992f). Chlorpyrifos is rapidly metabolized and excreted based on studies in
animals (Hayes and Laws, 1991).
5.4.1.3 Acute Toxicity—
Effects commonly associated with acute high-level exposure to Chlorpyrifos
include the following: headache, dizziness, weakness, incoordination, muscle
twitching, tremor, nausea, abdominal cramps, diarrhea, sweating, blurred or dark
vision, confusion, tightness in the chest, wheezing, productive cough, pulmonary
edema, slow heartbeat, salivation, tearing, toxic psychosis with manic or bizarre
behavior, influenza-like illness with weakness, anorexia, malaise, incontinence,
unconsciousness, and convulsions (HSDB, 1999).
5.4.1.4 Chronic Toxicity—
IRIS provides an oral RfD of 0.003 mg/kg-d based on a NOAEL in a 20-day study
reported in 1972 that found cholinesterase inhibition in adult male humans after
9 days of exposure. There were four subjects per dosed group. An uncertainty
factor of 10 was used to calculate the RfD (IRIS, 1999). There are limitations in
the use of this study for a chronic toxicity RfD. Although effects were observed
at levels lower than the NOAEL, they were discounted due to an inability to
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achieve statistical significance; however, it is very difficult to achieve statistical
significance with four subjects. No uncertainty factor was applied for the acute
nature of the study. Most important, EPA is reviewing its methods for evaluating
cholinesterase inhibitors. Cholinesterase inhibition alone is not necessarily
considered an adverse effect in the absence of other effects. The value listed on
IRIS was confirmed in 1993 by an Office of Pesticide Programs RfD Peer-Review
Committee (U.S. EPA, 1993c).
Other chronic exposure effects have been observed in study animals. In a 1991
two-generation rat study, adrenal lesions were reported at 1 and 5 mg/kg-d. In
a subchronic study at higher doses, the same effects were observed along with
increased brain and heart weight (U.S. EPA, 1992f).
There are significant uncertainties regarding an appropriate threshold for effects
of chlorpyrifos exposure. These include the very limited data on the recently
identified adrenal and cardiac effects of chlorpyrifos and the utility of a
cholinesterase endpoint. The IRIS value was used to calculate fish consumption
limits shown in Section 4 for chronic toxicity. Future improvements in the
database may result in alteration in this recommended value.
5.4.1.5 Reproductive and Developmental Toxicity—
Chlorpyrifos has been evaluated for developmental toxicity in mice, rats, and
rabbits (U.S. EPA, 2000b). Most studies only show effects at doses that cause
maternal toxicity due to cholinesterase inhibition (i.e., > 5 mg/kg-d). However, in
a study where observations were carried out to postnatal day 66, delayed
alterations in brain development were noted in offspring of rats receiving 1.0
mg/kg-d. Decreases in measurements of the parietal cortex were observed in
females. In addition, several studies show that neonates and young animals are
more susceptible to chlorpyrifos-induced cholinesterase inhibition than adults
(U.S. EPA, 2000b). For these reasons, a Population Adjusted Dose has been
calculated to provide extra protection for infants, children, and women of
childbearing age.
5.4.1.6 Mutagenicity—
Chlorpyrifos was not mutagenic in bacteria or mammalian cells. Slight genetic
alterations in yeast and DMA damage in bacteria have been observed.
Chlorpyrifos did not induce chromosome aberrations in vitro, was not clastogenic
in the mouse micronucleus test in vivo, and failed to induce unscheduled DMA
synthesis in isolated rat hepatocytes (U.S. EPA, 2000b).
5.4.1.7 Carcinogenicity—
Chlorpyrifos did not increase cancer incidence in 2-yr feeding studies in mice and
rats (U.S. EPA, 1992f). EPA has classified chlorpyrifos as Group E (evidence of
noncarcinogenicity for humans) (U.S. EPA, 1999c).
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5.4.1.8 Special Susceptibilities-
There is a recognized human population that may be at high risk with respect to
organophosphate exposure. Approximately 3 percent of the human population
has an abnormally low plasma cholinesterase level resulting from genetic causes.
These people are particularly vulnerable to cholinesterase-inhibiting pesticides.
Others at greater risk include persons with advanced liver disease, malnutrition,
chronic alcoholism, and dermatomyositis because they exhibit chronically low
plasma cholinesterase activities. Red blood cell (RBC) acetylcholinesterase is
reduced in certain conditions such as hemolytic anemias; people with these
conditions may be at greater risk than the general population from exposure to
organophosphates (U.S. EPA, 1999).
5.4.1.9 Interactive Effects-
No data were located. However, it is possible that coexposure to compounds with
a similar mechanism of action (i.e., organophosphate and carbamate pesticides)
may result in additive or synergistic effects.
5.4.1.10 Critical Data Gaps-
Data are needed on potential noncholinesterase effects of chronic exposure and
on the mechanism of toxicity that underlies the alterations in brain development
observed in the offspring of chlorpyrifos-treated rats. Additionally, toxicokinetic
data are needed to explain the differential extent of cholinesterase inhibition
between adult and young animals.
5.4.1.11 Summary of EPA Health Benchmarks-
Chronic Toxicity 3 x 10~4 mg/kg-d (3 x 10~5 mg/kg-d for infants, children, and
women ages 13-50)
Carcinogenicity Group E (evidence of noncarcinogenicity for humans).
5.4.1.12 Major Sources—
HSDB (1993), IRIS (1999), U.S. EPA (1992f, 2000b).
5.4.2 Diazinon
5.4.2.1 Background—
Diazinon is an organophosphorus insecticide that has been used widely since its
introduction in 1952. Most use is agricultural, although diazinon formulations are
also used commercially and by the general public for home, lawn, and garden
insect control.
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5.4.2.2 Pharmacokinetics—
Diazinon is converted in the liver into its active form diazoxon. Both diazinon and
diazoxon are rapidly deactivated by esterases in the blood and liver. Animal
studies indicate that diazinon and its metabolites are cleared from all tissues in the
body within 12 days after single exposures (ATSDR, 1996d). Human milk may
contain trace amounts of diazinon based on the results of exposure in cows
(HSDB, 1993).
5.4.2.3 Acute Toxicity—
Diazinon is highly toxic. The estimated adult oral fatal dose is approximately 25 g
(HSDB, 1993). Toxic effects are seen in the central and peripheral nervous
system due to inhibition of cholinesterase.
5.4.2.4 Chronic Toxicity—
There is currently no IRIS file for diazinon. However, OPP provides an RfD of
7 x 10~4 mg/kg-d based on a NOAEL of 0.025 mg/kg-d observed for plasma
cholinesterase inhibition in a human study. The uncertainty factor was 30:10 for
intraspecies variability and 3 for the protection of infants and children (U.S. EPA,
1998b).
Very little dose-response data are available on chronic systemic toxicity otherthan
cholinesterase effects. Hematocrit depression was observed in a rat chronic
feeding study at 50 mg/kg-d. Gastrointestinal disturbances were noted at 5
mg/kg-d with a NOEL of 0.05 mg/kg-d in a chronic monkey study (U.S. EPA,
1993d). If an alternative to cholinesterase inhibition is required, the monkey study
can be used with standard uncertainty factors that take into consideration inter-
and intraspecies variability.
5.4.2.5 Reproductive and Developmental Toxicity—
The reproductive/teratogenic studies listed in the tox one-liners report no adverse
effects at the highest doses tested (U.S. EPA, 1993d).
HSDB reported multiple studies indicating diazinon is teratogenic; however, the
relevance of these studies is questionable since they were not conducted using
standard protocols and administration of diazinon was by parenteral routes. In a
prenatal exposure study (dose not specified), multiple doses of diazinon resulted
in a higher incidence of urinary malformations, hydronephrosis, and hydroureter.
Diazinon was teratogenic in rats administered a single dose on day 11 of
gestation. Decreased fetal body weight was the most sensitive indicator. No
dose was specified in the database (HSDB, 1993). In chicks, diazinon exposure
led to abnormal vertebral column development including a tortuous and shortened
structure with abnormal vertebral bodies. In the neck region, the vertebral bodies
had fused neural arches and lacked most intervertebral joints. More severe
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effects on other elements of the skeleton were observed at higher doses (HSDB,
1993; Hayes, 1982). The dose (1 mg/egg) is not easily convertible to a
mammalian dose.
Behavioral effects were observed in mice exposed prenatally at 0.18 and 9 mg/kg-
d throughout gestation. The high-dose group showed decreased growth, several
behavioral effects, and structural pathology of the forebrain. The low-dose group
did not have brain pathology or growth abnormalities; however, they showed small
but measurable defects in behavior and a delay in reaching maturity (ATSDR,
1996d)
5.4.2.6 Mutagenicity—
Most mutagenicity assays were negative; one positive sister chromatid exchange
assay was noted (U.S. EPA, 1993d). A study on the effect of diazinon on mitosis
in human lymphocytes reported chromosomal aberrations in 74 percent of the
cells at 0.5 mg/mL (HSDB, 1993).
5.4.2.7 Carcinogenicity—
No evidence of carcinogenicity was observed in several long-term feeding studies
with diazinon in rodents (ATSDR, 1996d). EPA has classified diazinon as "not
likely" to be a human carcinogen (U.S. EPA, 1999c).
5.4.2.8 Special Susceptibilities-
There is a recognized human population that may be at high risk with respect to
organophosphate exposure. Approximately 3 percent of the human population
has an abnormally low plasma cholinesterase level resulting from genetic causes.
These people are particularly vulnerable to cholinesterase-inhibiting pesticides.
Others at greater risk include persons with advanced liver disease, malnutrition,
chronic alcoholism, and dermatomyositis because they exhibit chronically low
plasma cholinesterase activities. Red blood cell (RBC) acetylcholinesterase is
reduced in certain conditions such as hemolytic anemias; people with these
conditions may be at greater risk than the general population from exposure to
organophosphates (U.S. EPA, 1999).
5.4.2.9 Interactive Effects—
MIXTOX has reported antagonistic effects between diazinon and toxaphene with
exposure in rats via gavage (MIXTOX, 1992).
5.4.2.10 Critical Data Gaps—
OPP lists the following data gaps: reproduction study in rats, chronic feeding
oncogenicity study in rats, and chronic feeding study in dogs (U.S. EPA, 1992d).
A multigeneration reproductive study that evaluated developmental effects at low
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5.4 ORGANOPHOSPHATE PESTICIDES
doses and defined a NOAEL would be useful in establishing an appropriate RfD.
5.4.2.11 Summary of EPA Health Benchmarks-
Chronic Toxicity 7 x 10~4 mg/kg-d based on cholinesterase inhibition
Carcinogenicity "Not likely" to be a human carcinogen.
5.4.2.12 Major Sources—
ATSDR (1996d), Hayes (1982), HSDB (1993), U.S. EPA (1993d).
5.4.3 Disulfoton (Disyston)
5.4.3.1 Background—
Disulfoton is an organophosphate pesticide used on a wide variety of crops; major
uses are on corn, wheat, potatoes, and cotton. It is also used on fruit and nut
trees and ornamental plants.
5.4.3.2 Pharmacokinetics—
Disulfoton is readily absorbed after ingestion. Metabolism of disulfoton involves
sequential oxidation of the thioether sulfur and/or oxidative desulfuration in
addition to hydrolytic cleavage. The major metabolites are the sulfoxide acid
sulfone analogs of the compound. These are toxic metabolites that are degraded
rapidly to water-soluble nontoxic metabolites. Their estimated half-life is 30 to 32
hours (U.S. EPA, 1993f). Disulfoton is rapidly absorbed through the mucous
membrane of the digestive system and conveyed by the blood to body tissues.
The kidneys are the main route of elimination of the metabolites (HSDB, 1993).
5.4.3.3 Acute Toxicity—
The acute oral LD50 in animals ranges from 2 to 27.5 mg/kg (U.S. EPA, 1993f).
Disulfoton is highly toxic to all mammals by all routes of exposure (HSDB, 1993).
5.4.3.4 Chronic Toxicity—
IRIS provides an RfD of 4.0 x 10~5 mg/kg-d based on an LOAEL of 0.04 mg/kg-d
from a 2-year rat study that demonstrated cholinesterase inhibition and optic
nerve degeneration (IRIS, 1999). An uncertainty factor of 100 was used to
account for the interspecies differences and the spectrum of sensitivity in the
human population, plus a 10-fold factor to account for the lack of a no-effect level.
Numerous other effects of disulfoton have been reported at doses within 1 order
of magnitude of the LOAEL identified in the critical study. Toxicity, as reflected in
changes in absolute and relative organ weights, has been observed at 0.1 mg/kg-
d (the lowest dose tested) for the following systems: spleen, liver, pituitary, brain,
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5.4 ORGANOPHOSPHATE PESTICIDES
seminal vesicles, and kidneys (IRIS, 1999). In addition, at 0.65 mg/kg-d, rats
exhibited atrophy of the pancreas, chronic inflammation and hyperplasia in the
stomach, and skeletal muscle atrophy (U.S. EPA, 1993h).
5.4.3.5 Reproductive and Developmental Toxicity—
In a rat teratogenicity study, incomplete ossification of the parietals and
sternebrae were noted at 1 mg/kg-d with a NOEL of 0.3 mg/kg-d in rats. In a
1966 three-generation reproduction study in rats, male offspring had juvenile
hypoplasia in the testes, females had mild nephropathy in the kidneys, and both
had preliminary stages of liver damage at 0.5 mg/kg-d. No NOAEL was obtained,
and no data were provided on a number of critical parameters, including weight,
growth rate, and number of stillborn animals. Insufficient histologic data and
incomplete necropsy reports were identified by EPA reviewers (IRIS, 1999, U.S.
EPA, 1993f).
A more recent two-generation rat study identified a NOAEL of 0.04 mg/kg-d with
an LOAELof 0.12 mg/kg-d based on decreased litter sizes, pup survival, and pup
weights (U.S. EPA, 1993f).
5.4.3.6 Mutagenicity—
Disulfoton was not mutagenic in most assays; however, it was positive for
unscheduled DNA synthesis without activation in human fibroblasts, in a reverse
mutation assay in Salmonella (U.S. EPA, 1993f), and in other in vitro assays
(HSDB, 1993).
5.4.3.7 Carcinogenicity—
Insufficient information is available to determine the carcinogenic status of
disulfoton. Disulfoton has been classified as Group E (evidence of
noncarcinogenicity for humans (U.S. EPA, 1999c)
5.4.3.8 Special Susceptibilities-
Based on the organ toxicities observed in animal studies, individuals with
diseases or disorders of the following systems may be at greater risk from
exposure to disulfoton: pancreas, stomach, spleen, liver, pituitary, brain, seminal
vesicles, kidneys, musculoskeletal, and ocular. In addition, children who were
exposed prenatally to disulfoton may be at risk, depending on the level of
exposure.
5.4.3.9 Interactive Effects-
No data were located.
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5.4.3.10 Critical Data Gaps—
The IRIS file notes that additional rat reproduction studies and studies to evaluate
the ocular effects of disulfoton are needed (IRIS, 1999). HSDB notes that,
because of data gaps, a full risk assessment cannot be completed. Major relevant
data gaps noted under the Federal Insecticide Fungicide, and Rodenticide Act
(FIFRA) heading in HSDB include chronic toxicity, oncogenicity, and mutagenicity
data; animal metabolism; subchronic toxicity; and human dietary and nondietary
exposures (some data gaps may have been filled, cited in HSDB, 1993). As
noted above, additional studies are needed to identify the NOEL for sensitive
measures of the testicular, liver, and kidney toxicity identified in the
multigeneration study.
5.4.3.11 Summary of EPA Health Benchmarks-
Chronic Toxicity 4 x 10~5 mg/kg-d
Carcinogenicity Group E (evidence of noncarcinogenicity for humans).
5.4.3.12 Major Sources—
HSDB (1993), IRIS (1999), U.S. EPA (1993f).
5.4.4 Ethion
5.4.4.1 Background—
Ethion is an organophosphate pesticide used primarily on citrus crops (U.S. EPA,
1993a).
5.4.4.2 Pharmacokinetics—
Absorption of ethion is rapid by the oral route. Ethion is desulfurated by
P-450 enzymes in the liver to its active form, ethion monooxon, which causes
toxicity because of its potent inhibition of neural cholinesterase. Ethion and its
oxon form are detoxified by the action of esterases in the blood and liver,
producing diethyl phosphate and other metabolites that have not been
characterized. Ethion and its metabolites were cleared from the body within 7
days after single dose experiments in animals (ATSDR, 1998a).
5.4.4.3 Acute Toxicity—
Effects commonly associated with acute high-level exposure to ethion include the
following: headache, dizziness, weakness, incoordination, muscle twitching,
tremor, nausea, abdominal cramps, diarrhea, sweating, blurred or dark vision,
confusion, tightness in the chest, wheezing, productive cough, pulmonary edema,
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5.4 ORGANOPHOSPHATE PESTICIDES
slow heartbeat, salivation, tearing, toxic psychosis with manic or bizarre behavior,
influenza-like illness with weakness, anorexia, malaise, incontinence,
unconsciousness, and convulsions (HSDB, 1999).
5.4.4.4 Chronic Toxicity—
IRIS provides an RfD of 5 x 10~4 mg/kg-d based on two principal studies. A study
of 10 men reported a NOAEL of 0.05 mg/kg-d for plasma cholinesterase inhibition.
A second study in dogs reported a NOAEL of 0.06 and 0.07 mg/kg-day for males
and females, respectively, for plasma and brain cholinesterase inhibition
Uncertainty factors of 10 each were applied for intraspecies sensitivity and
because of concern for the significant effect on brain cholinesterase observed at
the next highest dose (0.71 mg/kg-d) in the dog study (IRIS, 1999, U.S. EPA,
1999d).
5.4.4.5 Reproductive and Developmental Toxicity—
In a rat developmental toxicity study, both the maternal and developmental toxicity
NOAELs were 0.6 mg/kg-d. Both the maternal and developmental toxicity
LOAELs were 2.5 mg/kg-d based on signs of hyperactivity in the parents. In a
rabbit developmental toxicity study, the NOAEL and LOAEL for maternal toxicity
were 2.4 and 9.6 mg/kg-d, respectively, based on weight loss, reduced food
consumption, and orange colored urine. The NOAEL for developmental toxicity
was 9.6 mg/kg-d, the highest dose tested (U.S. EPA, 1999d).
In a three-generation reproductive study in rats, the reproductive NOAEL was
1.25 mg/kg-d, the highest dose tested. The systemic toxicity NOAEL was 0.2
mg/kg-d and the LOAEL was 1.25 mg/kg-d based on decrease in serum
cholinesterase activity in P., and F2 female rats (U.S. EPA, 1999d).
5.4.4.6 Mutagenicity—
Ethion has shown no evidence of genotoxicity in several in vitro tests. Ethion was
negative in tests for point mutations, DNA repair, recombination, sister chromatid
exchange, and unscheduled DNA synthesis (ATSDR, 1998a). No in vivo tests of
ethion genotoxicity were located.
5.4.4.7 Carcinogenicity—
In 2-yr feeding studies with ethion in rodents, no evidence of carcinogenicity was
observed in rats (up to 2 mg/kg-d) or mice (up to 1.2 mg/kg-d) (ATSDR, 1998a).
On this basis, ethion is classified as a Group E chemical, evidence of non-
carcinogenicity for humans (U.S. EPA, 1999d).
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5.4.4.8 Special Susceptibilities—
In the case of ethion, EPA's Office of Pesticide Programs considered that a 10X
safety factor was not necessary for the protection of infants and children. This
recommendation was based on the following weight of evidence: no evidence of
enhanced susceptibility in fetuses in developmental studies in rats and rabbits, no
enhanced susceptibility in pups in a two-generation reproductive study in rats, no
evidence of developmental neurotoxicity, and completeness of the toxicology
database to assess susceptibility to infants and children (U.S. EPA, 1999).
5.4.4.9 Interactive Effects—
Potentiation between ethion and malathion has been observed. In rats, the
potentiation was approximately 2.9-fold. In dogs, there was very slight, if any,
potentiation (U.S. EPA, 19931).
5.4.4.10 Critical Data Gaps—
IRIS lists a chronic dog feeding study as a data gap (IRIS, 1999).
5.4.4.11 Summary of EPA Health Benchmarks-
Chronic Toxicity 5 x 10~4 mg/kg-d
Carcinogenicity Group E (evidence of noncarcinogenicity for humans).
5.4.4.12 Major Sources—
ATSDR (1998a), IRIS (1999), U.S. EPA (1999d).
5.4.5 Terbufos
5.4.5.1 Background—
Terbufos is an organophosphorus insecticide/nematicide applied to the soil to
control insects in a variety of crops.
5.4.5.2 Pharmacokinetics—
After a single oral dose of terbufos in rats, 83 percent was eliminated in urine as
metabolites and 3.5 percent in the feces over the following 7 days. No unusual
distribution of terbufos or its metabolites was noted in tissues (U.S. EPA, 1995).
5.4.5.3 Acute Toxicity—
Terbufos has a high acute toxicity to humans. Animal studies yielded the
following results: an oral LD50 in rats of 1.3 to 1.6 mg/kg (surveillance index) and
an oral LD50 in mice of 1.3 to 6.6 mg/kg (U.S. EPA, 1992e).
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5.4 ORGANOPHOSPHATE PESTICIDES
5.4.5.4 Chronic Toxicity—
Limited information is available on terbufos toxicity, and the focus of most toxicity
evaluations is on its cholinesterase inhibition properties. There is currently no
IRIS file for terbufos. The OPP lists an RfD of 2 x 10"5 mg/kg-d based on a
NOAEL of 0.005 mg/kg-d for plasma cholinesterase inhibition in a 28-day study
in dogs (U.S. EPA, 1997h). An uncertainty factor of 300 (10 for interspecies
variation, 10 for intraspecies variation, and 3 for protection of infants and children)
was applied to the NOAEL.
Quantitative chronic toxicity information on cholinesterase inhibition is available.
In rats, a 1974 lifetime oral study found a LOAEL of 0.0125 mg/kg-d (the lowest
dose tested); a 19871-year oral study found a NOAEL of 0.025 mg/kg-d. In dogs,
a 1972 6-month oral study found a NOAEL of 0.0025 mg/kg; a 1986 1-year study
found a LOAEL of 0.015 mg/kg-d (the lowest dose tested); a 1987 28-d dog study
identified a NOAEL of 0.00125 mg/kg-d (U.S. EPA, 1992e).
Quantitative data on chronic effects that are not directly related to cholinesterase
inhibition are limited because of the lack of "no effect levels" from many studies
and the need for specific information on some effects. Chronic exposure effects
include: corneal cloudiness and opacity, eye rupture, alopecia, disturbances in
balance, and exophthalmia noted in multiple studies and multiple species at
0.0125 mg/kg-d and above (U.S. EPA, 1992e). Increased liver weight and
increased liver extramedullary hematopoiesis at 0.025 mg/kg-d and above, and
mesenteric and mandibular lymph node hyperplasia at 0.05 mg/kg-d and above
were noted in a subchronic (3-mo) rat study (animals were not examined for this
lesion at lower exposure levels) (U.S. EPA, 1992e).
5.4.5.5 Reproductive and Developmental Toxicity—
Data currently available on developmental toxicity are limited because the
endpoints identified were gross measures of toxicity (death) and the underlying
causes of toxicity were not identified. The studies are not based on sensitive
measures of developmental toxicity. Results from two developmental studies and
one multigeneration study are available: a 1984 rat study found a NOAEL of 0.1
mg/kg-d with increased fetal resorptions at 0.2 mg/kg-d; a 1988 rabbit study
identified a NOAEL of 0.25 mg/kg-d with fetal resorptions at 0.5 mg/kg-d. A 1973
multigeneration reproductive study found a NOAEL of 0.0125 mg/kg-d in rats,
based on an increase in the percentage of deaths in offspring (U.S. EPA, 1992e).
5.4.5.6 Mutagenicity—
Terbufos was negative in most assays. It was positive in an in vivo dominant-
lethal assay in rats; at 0.4 mg/kg, the numbers of viable implants was reduced
(U.S. EPA, 1992e).
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5.4 ORGANOPHOSPHATE PESTICIDES
5.4.5.7 Carcinogenicity—
EPA has classified terbufos as Group E, evidence of noncarcinogenicity for
humans (U.S. EPA, 1999c).
5.4.5.8 Special Susceptibilities-
There is a recognized human population that may be at high risk with respect to
organophosphate exposure. Approximately 3 percent of the human population
has an abnormally low plasma cholinesterase level resulting from genetic causes.
These people are particularly vulnerable to cholinesterase-inhibiting pesticides.
Others at greater risk include persons with advanced liver disease, malnutrition,
chronic alcoholism, and dermatomyositis because they exhibit chronically low
plasma cholinesterase activities. Red blood cell (RBC) acetylcholinesterase is
reduced in certain conditions such as hemolytic anemias; people with these
conditions may be at greater risk than the general population from exposure to
organophosphates (U.S. EPA, 1999).
5.4.5.9 Interactive Effects-
No data were located.
5.4.5.10 Critical Data Gaps-
There are inconsistencies in the toxicity database for terbufos based on a
comparison of acute study results and the results obtained in some chronic
feeding studies, developmental studies, and the LD50s. Some longer-term studies
reported no effects at exposure levels above the LD50s (U.S. EPA, 1992e).
The animal and human studies available on terbufos do not provide a complete
and consistent basis for calculation of an alternative exposure limit. The identifi-
cation of mesenteric and mandibular lymph node hyperplasia is problematic due
to its potential oncogenic implications. A NOAEL for these effects was not
identified and effects were not screened in low-dose groups. Other effects, which
are not directly related to cholinesterase inhibition, were also noted with terbufos
exposure, including optic damage at 0.0125 mg/kg-d in multiple species and
studies. In addition, there is uncertainty regarding a safe exposure level to
prevent adverse developmental effects, as discussed above. These results
warrant further evaluation and may be considered, by some, to justify an
additional modifying factor to deal with data gaps and uncertainties in the
database.
5.4.5.11 Summary of EPA Health Benchmarks-
Chronic Toxicity 2 x 10~5 mg/kg-d.
Carcinogenicity Group E (evidence of noncarcinogenicity for humans).
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5.4 ORGANOPHOSPHATE PESTICIDES
5.4.5.12 Major Sources—
HSDB (1993), U.S. EPA (1992e).
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5.5 CHLOROPHENOXY HERBICIDES
5.5 CHLOROPHENOXY HERBICIDES
5.5.1 Oxyfluorfen
5.5.1.1 Background—
Oxyfluorfen is a recently introduced diphenyl ether pesticide in the chlorophenoxy
class. Limited data were located on this chemical.
5.5.1.2 Pharmacokinetics—
No data were located.
5.5.1.3 Acute Toxicity—
The acute oral LD50 in rats is greater than 5,000 mg/kg (Hayes and Laws, 1991).
5.5.1.4 Chronic Toxicity—
IRIS provides an RfDofS x 10~3mg/kg-d based on a NOAELof 0.3 mg/kg-d from
a 1977 20-month mouse feeding study that identified nonneoplastic lesions in the
liver and increased absolute liver weight. Uncertainty factors of 10 each for inter-
and intraspecies sensitivity were applied (IRIS, 1999).
5.5.1.5 Reproductive and Developmental Toxicity—
A three-generation rat study provided a NOAEL of 0.5 mg/kg-d and an LOAEL of
5 mg/kg-d. A rat teratology study identified a NOAEL of 100 mg/kg-d. A rabbit
study found fused sternebrae at 30 mg/kg-d and a NOEL of 10 mg/kg-d (IRIS,
1999, U.S. EPA, 1993J). A rabbit teratology study data gap is noted in the IRIS file
(IRIS, 1999).
5.5.1.6 Mutagenicity—
Results of mutagenicity assays on Oxyfluorfen are mixed (U.S. EPA, 1993J).
5.5.1.7 Carcinogenicity—
Oxyfluorfen has been classified as a possible human carcinogen (C) based on
liver tumors identified in experimental animals. A cancer slope factor of 0.0732
mg/kg-d has been derived (EPA 1998c).
5.5.1.8 Interactive Effects-
No data were located.
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5.5 CHLOROPHENOXY HERBICIDES
5.5.1.9 Critical Data Gaps—
The IRIS file notes a rabbit teratology study as a data gap (IRIS, 1999).
5.5.1.10 Summary of EPA Health Benchmarks-
Chronic Toxicity 3 x 10~3 mg/kg-d
Carcinogenicity 7.32 x 10~2 mg/kg-d.
5.5.1.11 Major Sources—
IRIS (1999), U.S. EPA(1993j).
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5.6 POLYCYCLIC AROMATIC HYDROCARBONS (PAHs)
5.6 POLYCYCLIC AROMATIC HYDROCARBONS (PAHs)
5.6.1 Background
Polycyclic aromatic hydrocarbons (PAHs) are a group of organic chemicals that
have a fused ring structure of two or more benzene rings. PAHs are also
commonly referred to as polynuclear aromatic hydrocarbons (PNAs). They are
formed during the incomplete combustion of organic materials. Industrial activities
that produce PAHs include coal coking; production of carbon blacks, creosote,
and coal tar; petroleum refining; synfuel production from coal; and the use of
Soderberg electrodes in aluminum smelters and ferrosilicum and iron works.
Domestic activities that produce PAHs include cigarette smoking, home heating
with wood or fossil fuels, waste incineration, broiling and smoking foods, and use
of internal combustion engines. PAHs are ubiquitous in the environment and
usually occur as mixtures. PAHs with two to five benzene rings are generally of
greatest concern for environmental and human health effects (U.S. EPA, 1999a).
ATSDR (1995b) has identified the following PAHs as the most important with
regard to human exposure:
• Acenaphthene
• Acenaphthylene
• Anthracene
• Benz[a]anthracene
• Benzo[a]pyrene
• Benzo[e]pyrene
• Benzo[Jb]fluoranthene
• Benzo[/c]fluoranthene
• Benzo[/]fluoranthene
• Benzo[g,/?,/]perylene
• Chrysene
• Dibenz[a,/?]anthracene
• Fluoranthene
• Fluorene
• lndeno[7,2,3-cd]pyrene
• Phenanthrene
• Pyrene.
Although these and many other PAHs are present in the environment,
benzo[a]pyrene is the chemical with most of the available health effects data.
5.6.2 Pharmacokinetics
PAHs may be absorbed through the lungs, the stomach, or the skin. The extent
of absorption varies in both humans and animals with the individual compound
and is influenced by vehicle. For instance, oral absorption increases with more
lipophilic PAHs or in the presence of oils in the intestinal tract. After inhalation,
oral, or dermal exposure of animals, the highest levels of PAHs were found in
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5.6 POLYCYCLIC AROMATIC HYDROCARBONS (PAHs)
highly perfused tissues, such as the lung, liver, gastrointestinal tract, and kidney.
Animal studies also show that PAHs cross the placenta. PAHs are rapidly
metabolized and excreted in humans and animals. The elimination half-life for
benzo[a]pyrene in rodents is 20 to 30 hours (ATSDR, 1995b).
PAHs have been shown (ATSDR, 1995b) to be metabolized to reactive
intermediates by enzyme systems commonly found in the lung, intestines, and
liver. These intermediates then covalently bind to cellular macromolecules,
leading to mutation and tumor development.
5.6.3 Acute Toxicity
Few data are available describing the acute toxicity of PAHs after inhalation
exposure in humans or animals. Limited information is available on the effects of
acute oral and dermal exposure in animals. However, benzo[a]pyrene is fatal to
mice following oral exposure to 120 mg/kg-d, and the liver and the skin have been
identified as target organs in animals after oral or dermal exposure, respectively
(ATSDR, 1998b).
5.6.4 Chronic Toxicity
Few controlled epidemiological studies have been reported in humans on the
effects of exposure to PAHs or to PAH-containing mixtures. However, available
information describing chronic-duration dermal exposure of humans to PAHs
indicates that PAHs have a high chronic exposure toxicity characterized by
chronic dermatitis and hyperkeratosis (ATSDR, 1995b).
Chronic studies in animals exposed to PAHs by ingestion, intratracheal
installation, or skin-painting have identified adverse effects on the cardiovascular,
respiratory, gastrointestinal, immune, and central nervous systems and on the
blood, liver, and skin (ATSDR, 1995b).
IRIS provides an RfD of 3 x 10"1 for anthracene based on a NOAEL of 1,000
mg/kg-d in a subchronic study in mice. Uncertainty factors of 10 each for inter-
and intraspecies variability were applied, with an additional uncertainty factor of
30 for use of a subchronic study and the lack of reproductive/developmental data
and adequate toxicity data in a second species. Confidence in the RfD is rated
low (IRIS, 1999).
An RfD of 4 x 10"2 mg/kg-d was calculated for fluoranthene based on a subchronic
study in mice, a NOAEL of 125 mg/kg-d, and critical effects on the liver, blood,
and kidneys. The same uncertainty factors were applied as for anthracene, with
confidence also rated as low (IRIS, 1999).
IRIS provides the same RfD for fluorene as for fluoranthene; a subchronic study
in mice was also used with the same NOAEL, uncertainty factors, and confidence
rating. For fluorene, the critical effect was on the blood (IRIS, 1999).
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5.6 POLYCYCLIC AROMATIC HYDROCARBONS (PAHs)
For pyrene, an RfD of 3 x 10~2 mg/kg-d was calculated. It was also based on a
subchronic study in mice, with a NOAEL of 75 mg/kg-d and kidney effects noted.
The same uncertainty factors and confidence rating were used (IRIS, 1999).
5.6.5 Reproductive and Developmental Toxicity
No information is available regarding the reproductive or developmental toxicity
of PAHs in humans. Animal data describing reproductive and developmental
effects of benzo[a]pyrene administered orally or parenterally and indicate that
PAHs have the potential to induce adverse reproductive and developmental
effects such as sterility, resorptions, and malformations (ATSDR, 1995b).
5.6.6 Mutagenicity
Benzo[a]pyrene has been thoroughly studied in genetic toxicology test systems
(ATSDR, 1995b). It induces genetic damage in prokaryotes, eukaryotes, and
mammalian cells in vitro and produces a wide range of genotoxic effects, including
gene mutations in somatic cells, chromosome damage in germinal and somatic
cells, DMA adduct formation, unscheduled DMA synthesis, sister chromatid
exchange, and neoplastic cell transformation. The genotoxic effects of the other
PAHs have been investigated using both in vivo and in vitro assays. All but three
of the PAHs (acenaphthene, acenaphthylene, and fluorene) were reported to be
mutagenic in at least one in vitro assay with the bacterium S. typhimurium.
5.6.7 Carcinogenicity
Evidence indicates that mixtures of PAHs are carcinogenic in humans. This
evidence comes primarily from occupational studies of workers exposed to
mixtures containing PAHs as a result of their involvement in such processes as
coke production, roofing, oil refining, or coal gasification (ATSDR, 1995b). Cancer
associated with exposure to PAH-containing mixtures in humans occurs
predominantly in the lung and skin following inhalation and dermal exposure,
respectively. In animals, individual PAHs have been shown to be carcinogenic by
the inhalation route (benzo[a]pyrene) and the oral route (e.g., benz[a]anthracene,
benzo[a]pyrene, and dibenz[a,/?]anthracene). Dermal exposure of animals to
benz[a]anthracene, benzo[a]pyrene, benzo[Jb]fluoranthene, benzo[/c]fluoranthene,
chrysene, dibenz[a,/?]anthracene, or indeno[7,2,3-cd]pyrene has been shown to
be tumorigenic in mice.
EPA has performed weight-of-evidence evaluations of several PAHs. The
carcinogenicity classifications are listed below (IRIS, 1999):
• Acenaphthylene D (not classifiable as to human carcinogenicity)
• Anthracene D
• Benz[a]anthracene B2 (probable human carcinogen)
• Benzo[a]pyrene B2
• Benzo[Jb]fluoranthene B2
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5.6 POLYCYCLIC AROMATIC HYDROCARBONS (PAHs)
• Benzo[/c]fluoranthene B2
• Benzo[g,/?,/]perylene D
• Chrysene B2
• Dibenz[a,/?]anthracene B2
• Fluoranthene D
• Fluorene D
• lndeno[7,2,3-cd]pyrene B2
• Phenanthrene D
• Pyrene D
EPA and others have developed a relative potency estimate approach for PAHs
(Nisbet and LaGoy, 1992; U.S. EPA, 1993n). Using this approach, the cancer
potency of the other carcinogenic PAHs can be estimated based on their relative
potency to benzo[a]pyrene. Table 5-2 lists the toxicity equivalence factors (based
on carcinogenicity) calculated by Nisbet and LaGoy (1992) for PAHs discussed
above.
U.S. EPA (1993n) has derived relative potency estimates based on mouse skin
carcinogenesis. These are shown in Table 5-3.
Table 5-2. Toxicity Equivalent Factors for Various PAHs
Compound Toxicity Equivalency Factor (TEF)
Dibenz[a,/?]anthracene 5
Benzo[a]pyrene 1
Benz[a]anthracene 0
Benzo[Jb]fluoranthene 0.1
Benzo[/<]fluoranthene 0.1
lndeno[7,2,3-cd]pyrene 0.1
Anthracene 0.01
Benzo[gA/]perylene 0.01
Chrysene 0.01
Acenaphthene 0.001
Acenaphthylene 0.001
Fluoranthene 0.001
Fluorene 0.001
Phenathrene 0.001
Pyrene 0.001
Source: Nisbet and LaGoy (1992).
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5.6 POLYCYCLIC AROMATIC HYDROCARBONS (PAHs)
Table 5-3. Relative Potency Estimates for Various PAHs
Compound Relative Potency3
Benzo[a]pyrene 1.0
Benz[a]anthracene 0.145
Benzo[Jb]fluoranthene 0.167
Benzo[/<]fluoranthene 0.020
Chrysene 0.0044
Dibenz[a,/?]anthracene 1.11
lndeno[7,2,3-cd]pyrene 0.055b
Source: U.S. EPA, 1993n.
a Model was P(d)=1 -exp[-a(1 +bd)2] for all but indeno[1,2,3-c,d]pyrene.
b Simple mean of relative potencies (0.021 and 0.089); the latter derived using
the one-hit model.
5.6.8 Special Susceptibilities
People with nutritional deficiencies, genetic diseases that influence the efficiency
of DMA repair, and immunodeficiency due to age or disease may be unusually
susceptible to the effect of PAHs. In addition, people who smoke, people with a
history of excessive sun exposure, people with liver and skin diseases, and
women, especially of reproductive age, may be at increased risk. Individuals with
hepatic-metabolizing enzymes that can be induced by PAHs may be unusually
susceptible to the toxic effects of PAH exposure by virtue of producing more toxic
metabolites. Fetuses may be susceptible to the effects of toxic PAH metabolites
produced by maternal exposure, because of increased permeability of the
embryonic and fetal blood-brain barrier and the immaturity of the enzymatic
systems that are responsible for elimination (ATSDR, 1995c).
5.6.9 Interactive Effects
Humans are usually exposed to PAHs in complex mixtures rather than to
individual PAHs. Interactions may occur among chemicals in a mixture prior to
exposure or may occur after exposure as a result of differing effects of the mixture
components on the body. Synergistic and/or antagonistic interactions with regard
to the development of health effects, particularly carcinogenesis, may occur. The
interaction between noncarcinogenicand carcinogenic PAHs have been examined
extensively in animals. Weakly carcinogenic or noncarcinogenic PAHs, including
benzo[e]pyrene, benzo[g,/?,/]perylene, fluoranthene, or pyrene exhibit co-
carcinogenic potential and tumor-initiating and promoting activity when applied
with benzo[a]pyrene to the skin of mice. In contrast, benzo[a]fluoranthene,
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5.6 POLYCYCLIC AROMATIC HYDROCARBONS (PAHs)
benzo[/c]fluoranthene, chrysene, and a mixture of anthracene, phenathracene,
and pyrene have been shown to significantly inhibit benzo[a]pyrene-induced
sarcoma after injection in mice. Several experiments have indicated that mixtures
of several PAHs are less potent with respect to carcinogenicity than the individual
PAHs that constitute the mixture (ATSDR, 1995c).
The majority of human exposure to PAHs occurs in the presence of particles or
other environmental pollutants that may influence the toxicity of the PAHs. For
instance, inhalation exposure to PAHs in the presence of particulate matter
greatly increases respiratory tract tumors in laboratory animals, due to the fact
that the particles are cleared more slowly from the lungs, thus allowing the
particle-bound PAHs to remain in the respiratory tract for longer periods of time.
Similarly, concomitant exposure to asbestos increases bronchopulmonary
cancers. Exposure to solvents or other environmental compounds that increase
metabolism of the PAHs may increase or decrease toxicity, depending on whether
the individual PAH must be transformed to toxic intermediates in order to exert its
adverse effect (ATSDR, 1995c).
5.6.10 Critical Data Gaps
A joint team of researchers from ATSDR, NTP, and EPA have identified the
following data gaps: human responses to acute, intermediate (14 to 365 days),
and chronic exposure, subchronic reproductive tests in various species,
developmental toxicity studies in two species, immunotoxicity studies of animals
and humans, and neurotoxicity studies in humans and animals (ATSDR, 1995c).
5.6.11 Summary of EPA Health Benchmarks
Chronic Toxicity (anthracene) 3 x 10~1 mg/kg-d
(fluoranthene) 4 x 10~2 mg/kg-d
(fluorene) 4 x 10~2 mg/kg-d
(pyrene) 3 x 10~2 mg/kg-d
Carcinogenicity (benzo[a]pyrene) 7.3 per mg/kg-d.
5.6.12 Major Sources
ATSDR (1998b), IRIS (1997c), U.S. EPA (1999a, 1993n).
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5.7 POLYCHLORINATED BIPHENYLS (PCBs)
5.7 POLYCHLORINATED BIPHENYLS (PCBs)
5.7.1 Background
Polychlorinated biphenyls (PCBs) are a mixture of chlorinated biphenyl chemicals
comprised of various chlorine substitution patterns. There are 209 possible PCB
congeners. Mixtures of PCBs were marketed in the United States under the trade
name Aroclor, with a numeric designation that indicated their chlorine content.
Although production and use of PCBs were banned in 1979, this chemical group
is extremely persistent in the environment and bioaccumulates through the food
chain. However, environmental mixtures of PCBs differ from the commercial
mixtures because of partitioning, transformation, and bioaccumulation. There is
evidence that some of the more toxic PCB congeners preferentially accumulate
in higher organisms (Aulerich et al., 1986). Consequently, the aggregate toxicity
of a PCB mixture may increase as it moves up the food chain (U.S. EPA, 1993a).
PCB exposure is associated with a wide array of adverse health effects in
experimental animals, but the effects of PCB exposure in humans are less clear.
Many effects have only recently been investigated (e.g., endocrine effects), and
the implications of newer studies are not fully known. The health effects of PCBs
are still under active evaluation and currently there is not sufficient information on
the specific congeners to develop congener-specific quantitative estimates of
health risk (ATSDR, 1998c; U.S. EPA, 1993a). Aroclor mixtures, rather than
environmental mixtures or bioconcentrated PCB mixtures, have been used in
laboratory animal studies to determine toxicity. The preferable studies would be
those that utilize human dose-response data from populations who have
consumed PCBs via fish or who have been exposed to PCBs in occupational
settings. Because sufficient human data are lacking, animal data were used to
develop RfDs and CSFs for PCBs. The Office of Water recommends that total
PCBs, calculated as the sum of the concentrations of the congeners or
homologue groups, be reported. Aroclor analysis is not recommended, except for
screening studies, because environmental PCB mixtures cannot be characterized
by any commercial Aroclor mixture (Cogliano, 1998). The first volume in this
document series, Sampling and Analysis, contains a detailed discussion of
analysis of this group of chemicals (U.S. EPA, 1993a).
5.7.2 Pharmacokinetics
PCBs are absorbed through the Gl tract and distributed throughout the body.
Studies of individual chlorobiphenyl congeners indicate, in general, that PCBs are
readily absorbed, with an oral absorption efficiency of 75 percent to greater than
90 percent (ATSDR, 1998b). Because of their lipophilic nature, PCBs, especially
the more highly chlorinated congeners (tetra- through hexachlorobiphenyl), tend
to accumulate in lipid-rich tissues. Greater relative amounts of PCBs are usually
found in the liver, adipose tissue, skin, and breast milk. It has been shown that
absorption of tetra- and higher chlorinated congeners from breast milk by nursing
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5.7 POLYCHLORINATED BIPHENYLS (PCBs)
infants ranges from 90 to 100 percent of the dose (ATSDR, 1998b). Offspring can
also be exposed to PCBs through placental transfer. PCBs have also been
measured in other body fluids including plasma, follicular fluid, and sperm fluid.
The retention of PCBs in fatty tissues is linked to the degree of chlorination and
also to the position of the chlorine atoms in the biphenyl ring. In general, higher
chlorinated PCBs persist for longer periods of time. Pharmacokinetics modeling
of PCB disposition indicates that PCB movement in the body is a dynamic
process, with exchanges between various tissues that depend on fluctuating
exposure levels to specific congeners. The result is elimination of congeners that
are more easily metabolized and retention of those that resist metabolism
(ATSDR, 1998c). In occupationally exposed individuals, lower chlorinated
congeners had half-lives between 1 and 6 years, whereas higher chlorinated
PCBs had half-lives ranging from 8 to 24 years (ATSDR, 1998b).
PCBs induce mixed function oxidases, and different congeners induce specific
forms (isozymes) of the cytochrome P-450 system. Although the mechanisms of
PCB toxicity have been investigated in many studies, a clear definition of the
mechanisms for most congeners has not been identified. The congeners appear
to act by a variety of mechanisms (ATSDR, 1998b). Some PCB congeners are
similar to dioxins and bind to a cytosolic protein, the Ah receptor, which regulates
the synthesis of a variety of proteins. The toxicity of these congeners is similar
to dioxins. The toxicity of other PCB congeners seems to be unrelated to the Ah
receptor. Ultimately, the toxicity of a PCB mixture depends on the toxicity of the
individual congeners, their interactions, and interactions with other chemical
contaminants such as pesticides and dioxins. For example, both synergistic and
antagonistic interactions have been reported with mixtures containing PCBs and
dioxins (Van den Berg et al., 1998).
5.7.3 Acute Toxicity
Acute high-level exposures of laboratory animals to PCBs have resulted in liver
and kidney damage, neurological effects, developmental effects,endocrine effects,
hematological effects, and death. LD50 values for various Aroclor mixtures range
from about 1,000 mg/kg to more than 4,000 mg/kg. No human deaths have been
associated with acute exposure to PCBs (ATSDR, 1998b).
5.7.4 Chronic Toxicity
In animal studies, numerous effects have been documented, including hepatic,
gastrointestinal, hematological, dermal, body weight changes, endocrine,
immunological, neurological, and reproductive effects (ATSDR, 1998b). Most of
the studies have involved oral exposure. Despite the variety of adverse effects
observed in animals exposed to PCBs, overt adverse effects in humans have
been difficult to document. This has been attributed to the fact that, in most
cases, the dosages tested in animals were considerably higher than those found
in occupational exposures and to the difficulties with interpreting epidemiological
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5.7 POLYCHLORINATED BIPHENYLS (PCBs)
studies (James et al.,1993; Kimbrough, 1995). These include multiple
confounding factors, uncertain exposure estimates, and statistical limitations.
Skin rashes and a persistent and severe form of acne (chloracne) have been
reported following exposures to PCBs. Occupational and accidental exposures
have indicated that PCBs may affect many organs including the gastrointestinal,
respiratory, immune, central nervous, and cardiovascular systems.
EPA has derived an RfD of 2 x m5 mg/kg-d for Aroclor 1254 (IRIS, 1999). The
RfD was based on a LOAEL of 0.005 mg/kg-d for ocular and immunological
effects in monkeys. The study reported ocular exudate and inflamed Meibomian
glands, distorted growth of finger and toenails, and decreased antibody response
(IgM and IgG) to injected sheep red blood cells at the lowest dose tested.
Uncertainty factors of 10 for sensitive individuals, 3 for extrapolation from
monkeys to humans, 3 for extrapolation from a subchronic exposure to a chronic
RfD, and 3 for use of a minimal LOAEL were applied, resulting in a total
uncertainty factor of 300. An uncertainty factory of 3 (rather than 10) for
extrapolation from subchronic to chronic exposure was used, because the
duration of the critical study continued for approximately 25 percent of the lifespan
of monkeys, and the immunologic and clinical changes observed did not appear
to be dependent upon duration.
EPA has medium confidence in the study used as the basis for the RfD for Aroclor
1254, in the database, and in the RfD. EPA based this rating on the fact that the
database consisted of a large number of laboratory animal and human studies;
however, there were some inconsistencies in the effect levels for reproductive
toxicity and the results of an unpublished study were considered (IRIS, 1999).
5.7.5 Developmental Toxicity
PCB mixtures have been shown to cause adverse developmental effects in
experimental animals (ATSDR, 1998c). Some human studies have also
suggested that PCB exposure may cause adverse effects in children and in
developing fetuses while other studies have not shown effects (U.S. EPA, 1999a).
Reported effects include lower IQ scores (Jacobson and Jacobson, 1996), low
birth weight (Rylander et al., 1998), and lower behavior assessment scores
(Lonky et al., 1996). However, study limitations, including lack of control for
confounding variables, and deficiencies in the general areas of exposure
assessment, selection of exposed and control subjects, and the comparability of
exposed and control samples. Different findings from different studies provide
inconclusive evidence that PCBs cause developmental effects in humans
(ATSDR, 1998b).
The RfD for Aroclor 1016 is based on reduced birth weights observed in monkeys
in a 22-month study (discussed below under longer-term developmental studies).
This study established a NOAEL of 0.007 mg/kg-d. Applying an uncertainty factor
of 100 (3 for sensitive individuals [infants exposed transplacentally], 3 for
interspecies extrapolation, 3 for database limitations [male reproductive effects
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5.7 POLYCHLORINATED BIPHENYLS (PCBs)
are not directly addressed and two-generation reproductive studies are not
available], and 3 for extrapolation from subchronic to chronic) to the NOAEL yields
anRfDof7x 10"5mg/kg-d (IRIS, 1999). However, since the RfD for Aroclor 1254
is more conservative (2 x 10~5 mg/kg-d) and protects against adult toxicity
concerns as well as the risk to the fetus and children, this RfD will be used to
calculate the consumption limits for all populations (adults, women of reproductive
age, and children).
EPA has medium confidence in the study, in the database, and in the RfD for
Aroclor 1016. EPA based this rating on the fact that the critical study was well
conducted in a sensitive animal species and the database for PCBs in general is
extensive; however, since mixtures of PCBs found in the environment do not
match the pattern of congeners found in Aroclor 1016, EPA felt that only a
medium confidence ranking could be given. For those particular environmental
applications where it is known that Aroclor 1016 is the only form of PCB
contamination, EPA stated that the RfD could be considered to have a high
confidence rating (IRIS, 1999).
A study was conducted of pregnancy outcomes in women who had consumed
PCB-contaminated fish from Lake Michigan over an average of 16 years
(exposure both prior to and during pregnancy). Consumption of contaminated fish
and levels of total PCBs in cord serum correlated with lower birth weight, smaller
head circumference, and shorter gestational age. Fish consumption, however,
was correlated with delayed neuromuscular maturity, and, at 7 months, the
children had subnormal visual recognition memory. Children from this cohort
were examined at age 4 and 11 years. At age 4, cord serum PCB levels were
associated with impaired short-term memory. Activity level was inversely related
to 4-year serum PCB level and also to maternal milk PCB level. At age 11,
prenatal exposure to PCBs was associated with lower full-scale and verbal IQ
scores after controlling for potential confounding variables, such as
socioeconomic status. The results from this series of studies were confounded
by possible maternal exposure to other chemicals and by the fact that the
exposed group, on average, drank more alcohol and caffeine, prior to and during
pregnancy, weighed more, and took more cold medications during pregnancy,
than the nonexposed group (Fein et al., 1984a, 1984b).
Other relevant studies generally found no significant differences between control
groups and exposed groups concerning stillbirths, multiple births, preterm births,
congenital anomalies, and low birth weight.
Information on chronic developmental toxicity is available from studies in Rhesus
monkeys (ATSDR, 1998b). Exposure periods ranged from 12 to 72 months.
Inflammation of tarsal glands, nail lesions, and gum recession were noted in
offspring of monkeys exposed to Aroclor 1254. Adverse neurobehavioral effects
were reported following exposure to Aroclor 1016 and Aroclor 1248. Other
observed effects included reduction in birth weight and increased infant death for
Aroclor 1248.
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5.7 POLYCHLORINATED BIPHENYLS (PCBs)
Exposure via lactation is a significant concern for neonates because PCBs
concentrate in milk fat. Animal studies indicate that lactational exposure, in some
cases, can be more significant than transplacental transfer. In monkeys, signs of
intoxication have been observed in offspring exposed to PCBs in maternal milk
(ATSDR, 1998b).
In summary, the results from some studies in humans suggest that exposure to
PCBs may cause developmental effects. However, limitations of these studies
diminished the validity of the results. Animal studies indicate that PCBs can
cause some developmental effects following prenatal or postnatal exposure.
5.7.6 Mutagenicity
The majority of mutagenicity assays of PCBs have been negative (IRIS, 1999).
However, an increase in the percentage of chromosomal aberrations in peripheral
lymphocytes and an increase in the sister chromatid exchange rate were reported
in a study of workers manufacturing PCBs for 10 to 25 years. Although workers
and controls were matched for smoking and drinking, concurrent exposure to
other known human genotoxic chemicals occurred (ATSDR, 1998b). Another
study found an increased incidence of chromatid exchanges in lymphocytes from
workers exposed to PCBs in an electric power substation fire compared to
unexposed controls. It is possibile that toxic chlorinated dioxins and/or furans
generated during the fire may have been responsible for the effects.
The weight of evidence from the in vitro and in vivo genotoxicity studies suggests
that PCBs are not likely to be genotoxic to humans. However, exposure to PCBs
may enhance the genotoxic activity of other chemicals (ATSDR, 1998b).
5.7.7 Carcinogenicity
PCBs are classified by EPA as Group B2; probable human carcinogens . This is
based on studies that have found liver tumors in rats exposed to Aroclors 1260,
1254, 1242, and 1016. Evaluation of the animal data indicate that PCBs with 54
percent chlorine content induces a higher yield of liver tumors in rats than other
PCB mixtures.
Human epidemiological studies of PCBs have not yielded conclusive results
(Silberhorn et al., 1990). There is some suggestive evidence that xenoestrogens,
including PCBs, may play a role in breast cancer induction (ATSDR, 1998c).
Some studies have indicated an excess risk of several cancers, including: liver,
biliary tract, gallbladder, gastrointestinal tract, pancreas, melanoma, and non-
Hodgkin's lymphoma (IRIS, 1999, ATSDR, 1998c). As with all epidemiological
studies, it is very difficult to obtain unequivocal results because of the long latency
period required for cancer induction and the multiple confounders arising from
concurrent exposures, lifestyle differences, and other factors. The currently
available human evidence is considered inadequate but suggestive that PCBs
may cause cancer in humans (IRIS, 1999).
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5.7 POLYCHLORINATED BIPHENYLS (PCBs)
The Agency's recent peer-reviewed reassessment published in a final report,
PCBs: Cancer Dose-Response Assessment and Application to Environmental
Mixtures (U.S. EPA, 1996c), adopts an innovative approach that distinguishes
among PCB mixtures by using information on environmental processes. It
considers all cancer studies (which used commercial mixtures only) to develop a
range of cancer slope factors, then uses information on environmental processes
to provide guidance on choosing an appropriate slope factor for representative
classes of environmental mixtures and different pathways. Depending on the
specific application, either central estimates or upper bound estimates can be
appropriate. Central estimates describe a typical individual's risk, while upper
bounds provide greater assurance that the true risk is not likely to be
underestimated. Central estimates are used for comparing or ranking
environmental hazards, while upper bounds provide information about the
precision of the comparison or ranking. In this reassessment, the use of the
upper bound values was found to increase cancer potency estimates by only two-
or threefold over those using central tendency. Upper bounds are useful for
estimating risks or setting exposure-related standards to protect public health and
are used by EPA in quantitative cancer risk assessment. Thus, the cancer
potency of PCB mixtures is determined using a tiered approach based on
environmental exposure routes with upper-bound slope factors ranging from 0.07
to 2 per mg/kg-d for average lifetime exposures to PCBs. It is noteworthy that
bioaccumulated PCBs appear to be more toxicthan commercial PCBs and appear
to be more persistent in the body (IRIS, 1999). In addition, there is evidence that
early-life exposures may result in an increased risk (U.S. EPA, 1996c). Therefore,
the highest cancer slope factor is recommended for the following conditions: food
chain exposure; sediment and soil ingestion; inhalation of dust or aerosols; dermal
exposure (if an absorption factor has been applied); presence of dioxin-like,
tumor- promoting, or persistent congeners; and early-life exposure.
Alternatively, if site-specific congener concentrations are available, the risk
assessment can be supplemented by determining the dioxin-like toxicity (U.S.
EPA, 1996c; Cogliano, 1998). Cogliano(1998) presents data showing the typical
composition of several commercial Aroclor mixtures (Table 5-4). Aroclors 1016,
1242, 1254, and 1260 contained concentrations of dioxin-like concentrations
ranging from 0.14 ppm to 46.4 ppm TEQs. Therefore, separate risk assessments
should be conducted for the dioxin-like and nondioxin-like PCB congeners if the
congener analysis indicates elevated concentrations of dioxin-like congeners
relative to the typical commercial mixtures (IRIS, 1999; U.S. EPA, 1996c).
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5.7 POLYCHLORINATED BIPHENYLS (PCBs)
Table 5-4. Reported Concentrations (ppm) of Dioxin-Like Congeners in
Commercial Aroclor Mixtures
Aroclors
Congener
PCB-77 (3,3',4,4'-tetraCB)
PCB-126 (3,3,4,4',5-pentaCB)
PCB-169 (3,3,4,4',5,51-hexaCB)
PCDFs
TEQ from PCBs
TEQ from PCDFs
1016
66
0.95
0.0
0.05
0.14
0.002
1242
3340
44
0.0
2.2
8.1
0.1
1254
918
134.3
1.52
0.13
46.4
0.01
1260
31
0.0
0.0
5.5
7.1
0.08
Source: Cogliano, 1998.
CB = Chlorinated biphenyls
In a recent study conducted by the Delaware Department of Natural Resources
and Environmental Control (Greene, 1999), dioxin-like PCBs, nondioxin-like
PCBs, and dioxins/furans accounted for about 64.4, 26.9, and 5.6 percent of the
total cancer risk, respectively, from ingesting fish caught from the Chesapeake
and Delaware Canal. Data from this study are shown in Table 5-5 to illustrate the
potential importance of the dioxin-like PCB congeners. The DDNREC noted that,
had cancer risk been calculated according to the traditional method (i.e., not
including a separate assessment for dioxin-like PCBs), the cancer risk estimate
for PCBs would have been lower by a factor of 2.9. However, PCBs contributed
about 93 percent of the total dioxin risk based on 2,3,7,8-TCDD TEQs. Therefore,
failure to evaluate the dioxin-like PCB congeners could result in underestimating
cancer risk.
Table 5-5. PCB and Dioxin Concentrations (ppb) in Channel Catfish
Parameter
Total PCBs
Nondioxin-like PCBs
Dioxin-like PCBs TEQs
Dioxin/furan TEQs
Total TEQs
Median
1,104.8
943.8
0.0302
0.0026
0.0328
Mean
1,173
1,024.9
0.0303
0.0024
0.0327
Maximum
1,665.3
1,474.7
0.0509
0.0043
0.0552
Source: Greene, 1999.
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5.7 POLYCHLORINATED BIPHENYLS (PCBs)
5.7.8 Special Susceptibilities
There is evidence that embryos, fetuses, and neonates are more susceptible to
PCBs due to their underdeveloped enzymatic systems, which may lead to
increased PCB accumulation in the body. Breast-fed infants may have an
increased risk because of bioconcentration of PCBs in breast milk and high intake
rates relative to body weights. In addition, there is evidence that a steroid present
in human milk inhibits glucuronyl transferase activity, which could, in turn, inhibit
glucuronidation and excretion of PCB metabolities. Other individuals with
potentially greater risk include those with liver and blood diseases or those with
syndromes associated with impairment to the metabolic systems that help
eliminate PCBs from the body.
5.7.9 Interactive Effects
PCBs induce microsomal enzymes; therefore, the effects of exposure to PCBs or
other compounds depends on the role of oxidative metabolism. For example,
preexposure to PCBs may enhance the liver toxicity of some chemicals
(trichloroethylene, mirex, kepone, carbon tetrachloride, tetrachloroethylene) but
decrease the liver toxicity of 1,1-dichloroethylene. Other interactive effects
include increased metabolism and excretion of pentobarbital, increased
genotoxicity of numerous carcinogens, increased duodenal ulcerogenic activity of
acrylonitrile, and increased kidney toxicity of trichloroethylene (ATSDR, 1998b).
5.7.10 Critical Data Gaps
The following studies could help fill in some of the key data gaps for PCBs:
congener-specific PCB levels in human tissues; epidemiological studies of
populations living near PCB-contaminated sites and occupational settings where
exposure to PCBs still occurs; reproductive studies in humans and animals,
including fertility studies in males of a sensitive species; developmental and
neurodevelopmental studies, immunotoxicity studies in humans and animals;
neurotoxicity studies in humans with high PCB body burdens and in animals;
chronic studies to determine the most sensitive animal target organ and species;
and comparative toxicity of Aroclors and bioaccumulated PCBs (ATSDR, 1998b).
5.7.11 Summary of EPA Health Benchmarks
Chronic Toxicity 2 x 10~5 mg/kg-d based on Aroclor 1254
Carcinogenicity 2.0 per mg/kg-d based on mixed PCBs.
5.7.12 Major Sources
ATSDR (1998b), Cogliano (1998), HSDB (1993), IRIS (1999), James et al.
(1993), Kimbrough (1995), Silberhorn etal. (1990), U.S. EPA (1996c).
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5.8 DIOXINS
5.8 DIOXINS
5.8.1 Background
Dioxins are a group of synthetic organic chemicals that contain 210 structurally
related individual chlorinated dibenzo-p-dioxins (CDDs) and chlorinated
dibenzofurans (CDFs). Dioxin is a generic term that is used, in this case, to refer
to the aggregate of all CDDs and CDFs. It is recommended that the 17 2,3,7,8-
substituted tetra- through octa-chlorinated dibenzo-p-dioxins and dibenzofurans
be considered together as a simplifying and interim approach until further
guidance is available on this chemical group. In addition, 12 PCB congeners have
been identified that exhibit dioxin-like activity (U.S. EPA, 1996c, Van den Berg et
al., 1998). The reader may consult guidance on the use of a toxicity equivalency
approach to refine the toxicity estimate and fish consumption limit calculations
(Barnes and Bellin, 1989; U.S. EPA, 1991c; U.S. EPA, 1996c).
Dioxin has been undergoing extensive review within EPA for several years.
Consequently, only a brief summary, is provided below. Currently, the EPA's
dioxin reassessment document, which includes two reports entitled Estimating
Exposure to Dioxin-like Compounds (three volumes) (U.S. EPA, 1994a) and
Health Assessment Document for 2,3,7,8-Tetrachlorodibenzo-p-dioxin (TCDD)
and Related Compounds (three volumes) (U.S. EPA, 1994b) is undergoing final
review. The dioxin reassessment document is scheduled for final external peer
review during the third quarter of fiscal year 2000. Following peer review, the
document will be sent to the EPA Science Advisory Board for final review. The
final dioxin reassessment document is scheduled for release by the end of the
calendar year 2000.
5.8.2 Pharmacokinetics—
Dioxins are absorbed through the gastrointestinal tract, respiratory tract, and skin
and distributed throughout the body. Absorption is congener-specific with
decreased absorption of hepta- and octa-congeners compared to dioxins with
fewer chlorines. Because of their lipophilic nature, dioxins tend to accumulate in
fat and the liver. Dioxins are slowly metabolized by oxidation or reductive
dechlorination and conjugation and the major routes of excretion are the bile and
feces. Reported half-lives in the body range from 5 to 15 years. Small amounts
may be eliminated in the urine. The current evidence indicates that metabolities
are less toxic than the parent compounds (ATSDR, 1998c, U.S. EPA 1994a).
The predominant forms retained in the tissues are the 2,3,7,8-substituted
congeners. Tissue deposition depends on the route of exposure, congeners
present, dose, and age. Based on a study of a human volunteer, about 87
percent of a single dose of dioxins dissolved in corn oil was absorbed and about
90 percent of the absorbed dose was distributed to fatty tissue (ATSDR, 1998c).
5-102
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5.8 DIOXINS
The half-lives for various dioxin congeners in humans have been reported to
range from 2.9 to 26.9 years. Some studies have suggested longer half-lives in
individuals with higher body fat (ATSDR, 1998c).
Dioxins induce mixed function oxidases and hepatic aryl hydrocarbon hydroxylase
(AHH). Dioxins bind to a cytosolic protein, the Ah receptor, which regulates the
synthesis of a variety of proteins. The Ah receptor has been found in many
human tissues, including the lung, liver, placenta, and lymphocytes. Although
evidence indicates that the Ah receptor is involved in many biological response
to dioxins, the diversity of biological effects observed cannot be accounted for by
characteristics of this receptor alone (ATSDR, 1998c, U.S. EPA, 1994a).
5.8.3 Acute Toxicity
LD50 values for dioxins vary over several orders of magnitude, depending on the
congener, species, and strain of animal tested. The most toxic congener is
2,3,7,8-TCDD with LD50 values ranging from 22 ,ug/kg to 340 ,ug/kg in various
strains of laboratory rats. Guinea pigs are the most sensitive species tested
(LD50s from 0.6 to 2.1 M9/kg) and hamsters are the most resistant (LD50s from
1,157 to 5,051 ,ug/kg). In all studies, the animals died from a pronounced wasting
syndrome characterized by weight loss and depletion of body fat that lasted 1 to
6 weeks. By contrast, laboratory animals have survived acute doses of 1 to 4
g/kg of 2,7-DCDD and OCDD. Single exposures to dioxins have also affected the
heart, liver, kidneys, blood, stomach, and endocrine systems of laboratory
animals. No human deaths have been directly associated with exposure to
dioxins. (ATSDR, 1998c).
5.8.4 Chronic Toxicity
In animal studies, numerous effects have been documented, including hepatic,
gastrointestinal, hematological, dermal, body weight changes, endocrine,
immunological, neurological, reproductive, and developmental effects. Most of
the studies have involved oral exposure. Despite the variety of adverse effects
observed in animals exposed to dioxins, adverse health effects in humans have
generally been limited to highly exposed populations in industrial factories or
following chemical accidents and contamination episodes. The adverse human
health effect most commonly associated with high-level exposure to dioxin-like
agents is the skin disease chloracne, a particularly severe and prolonged acne-
like skin disorder. Adverse human health effects were also noted following
consumption of heated rice oil contaminated with PCBs and CDFs. Conclusive
evidence of other adverse human health effects at lower dioxin exposure levels
is generally lacking because of incomplete exposure data, concomitant exposure
to other compounds, and/or small numbers of study participants. Some
epidemiological studies have suggested that dioxins may cause immuno-
suppression, respiratory effects, cardiovascular effects, and liver effects in
humans (ATSDR, 1998c, U.S. EPA, 1994a).
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5.8 DIOXINS
5.8.5 Reproductive and Developmental Toxicity
Dioxins have been shown to cause adverse developmental effects in fish, birds,
and mammals at low exposure levels. Several studies in humans have suggested
that dioxin exposure may cause adverse effects in children and in the developing
fetus. These include effects on the skin, nails, and meibomian glands;
psychomotor delay; and growth retardation. However, study limitations, including
lack of control for confounding variables, and deficiencies in the general areas of
exposure make it difficult to interpret these results. Overall, the human data are
inconclusive; however, the animal data suggest that developmental toxicity is a
concern (ATSDR, 1998c, U.S. EPA, 1994a).
In mammals, learning behavior and development of the reproductive system
appear to be among the most sensitive effects following prenatal exposure. In
general, the embryo or fetus is more sensitive than the adult to dioxin-induced
mortality across all species (ATSDR, 1998c, U.S. EPA, 1994a).
5.8.6 Mutagenicity
The majority of mutagenicity assays of dioxins have been negative. An increased
incidence of chromosomal aberrations was found in fetal tissue but not maternal
tissue in a group of women exposed to dioxins following an industrial accident in
Italy; however, cases treated for chloracne did not have an increased incidence
of chromosomal aberrations. Animal studies also are inconclusive. The available
data do not provide strong evidence that dioxins are genotoxic (ATSDR. 1998c,
U.S. EPA, 1994a).
5.8.7 Carcinogenicity
Dioxins are classified by EPA as Group B2 (sufficient evidence in animals,
insufficient evidence in humans) when considered alone and Group B1 (sufficient
evidence in animals, limited evidence in humans) when considered in association
with chlorophenols and phenoxyherbicides. This is based on studies that have
found multiple-site sarcomas and carcinomas in rats and mice exposed to various
dioxin mixtures and congeners. Epidemiological studies suggest an increased
incidence of cancer mortality (all types of cancers combined) and of some specific
cancers (soft-tissue sarcoma, non-Hodgkin's lymphoma, respiratory tract cancer,
and gastrointestinal cancers). In addition, there is evidence that 2,3,7,8-TCDD
acts as a tumor promoter. As with all epidemiological studies, it is very difficult to
obtain clear unequivocal results because of the long latency period required for
cancer induction and the multiple confounders arising from concurrent exposures,
lifestyle differences, and other factors. The currently available evidence suggests
that dioxins may cause cancer in humans (ATSDR, 1998c, U.S. EPA, 1994a).
EPA has derived a cancer slope factor of 1.56 x 105 (mg/kg-d)"1 for 2,3,7,8-TCDD
(HEAST, 1997).
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5.8 DIOXINS
5.8.8 Special Susceptibilities
There is evidence that children are more susceptible than adults to the dermal
toxicityofdioxins. Animal data suggest that the developing reproductive, immune,
and nervous systems of the fetus are particularly sensitive to dioxin toxicity
(ATSDR, 1998c).
5.8.9 Interactive Effects
Environmental exposure to dioxins includes various mixtures of CDDs, CDFs, and
some PCBs. These mixtures of dioxin-like chemicals cause multiple effects that
vary according to species susceptibility, congeners present, and interactions.
Risk assessment of these complex mixtures is based on the assumption that
effects are additive and there is some experimental evidence to support this.
However, there also is evidence that some interactions may result in inhibition and
others result in potentiation. Cotreatment of mice with various commercial PCB
mixtures (Aroclors) and 2,3,7,8-TCDD has resulted in inhibiting some of the Ah
receptor mediated responses. An increased incidence of cleft palate was
reported when mice were treated with both 2,3,7,8-TCDD and hexachlorobiphenyl
compared to treatment with 2,3,7,8-TCDD alone. Both synergistic and
antagonistic responses have been observed following co-exposure of 2,3,7,8-
TCDD with other chemicals as well (ATSDR, 1998c).
5.8.10 Critical Data Gaps
The following data gaps have been identified for dioxins: inhalation and dermal
toxicity studies; toxicity studies of dioxin compounds other than 2,3,7,8-TCDD;
continued medical surveillance of individuals with known past high exposures to
dioxins; mechanistic studies; immune function tests in human cohorts;
neurological tests in ongoing prospective studies of humans; congener-specific
human toxicokinetic studies to better assess human dosimetry; and further studies
to identify potential biomarkers for exposure and effects. Another critical data gap
is the need to gather exposure data and conduct modeling for the purpose of
linking human exposure to sources (ATSDR, 1998c).
5.8.11 Summary of EPA Health Benchmarks
Chronic Toxicity Not available
Carcinogenicity 1.56 x 10+5 per mg/kg-d.
5.8.12 Major Sources
U.S. EPA (1994a), ATSDR (1998c), Heast (1997), Van den Berg et.al. (1998).
5-105
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6. MAPPING TOOLS
SECTION 6
MAPPING TOOLS FOR RISK ASSESSMENT AND RISK MANAGEMENT
6.1 OVERVIEW OF POPULATION AND CONTAMINANT MAPPING
Mapping is useful for displaying geographic data concerning chemical con-
taminants, consumer populations, risks, locations of consumption advisories, or
other related information. Mapping allows risk assessors and risk managers to
work with a visual display of data that is easily understood and that may show
patterns of contamination and risk useful to risk managers. A variety of methods
for using mapping in risk assessment and management are discussed in this
section. Although presented in the risk assessment volume in this series, this
information may be useful to state staff in planning and displaying sampling and
analysis activities and results, as well as for risk management and risk com-
munication. Additional assistance with mapping may be obtained from mapping
software companies, university geography departments, and EPA Regional and
Headquarters offices that often use geographic information systems (GISs).
6.2 OBJECTIVES OF MAPPING
Mapping can be useful at every stage in the fish advisory development process
to
• Display sampling results with respect to fish species and chemical
contaminant levels
• Display population and/or fisher population density
• Display locations of recreational and subsistence fish harvests
• Spatially locate populations at high risk, based on high fish consumption rates
• Delineate areas where fish consumption advisories have been issued
• Determine where data gaps exist for purposes of targeting data collection
efforts appropriately.
Information can be mapped in various combinations to address specific concerns.
For example, mapping information on fisher population density and on
contaminant concentrations can be combined to produce an overview of
populations that may be at risk. Risk managers may find particular use for maps
showing locations where contamination exceeds screening levels or where a set
risk level is estimated to occur (e.g., greater than 100 percent of the RfD for
noncarcinogenic effects, greater than 1 in 1 million risk for carcinogens).
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6. MAPPING TOOLS
6.3 BASIC CIS CONCEPTS FOR POPULATION AND CONTAMINANT MAPPING
A GIS stores information about the world as a collection of thematic layers that
can be linked together by geography. A GIS is commonly defined as a computer
system designed to allow users to collect, manage, and analyze large volumes of
spatially referenced files and associated data layers. GISs are used for solving
complex research, planning, and management problems. The major components
of a GIS are: a computer with software providing a special user interface designed
to facilitate dealing with spatial databases (or layers); database management
software that allows spatial data sets to be created and maintained, along with
features for importing data from other computer systems; a set of software tools
to carry out spatial data processing and analyses of the GIS layers; and a high-
resolution display system (usually a graphics monitor and a high-quality printer or
plotter) to create the maps that summarize the spatial analysis work.
Two technologies have been developed for taking information about features in
the real world and converting these into a GIS data layer. Raster technologies
were developed largely in working with satellite images, high-altitude aerial
photographs, or other remote sensing data where the information is organized
around small squares or pixels similar to the "dots" found in the photographs
printed in books or newspapers. Vector technologies involve a richer set of
objects for breaking down the real world into features. Instead of small pixel
patches, vector technologies can organize data using a more intuitive set of
polygons (e.g., the boundary of a town), lines or arcs (e.g., rivers or roads), and
points (e.g., the location of aSuperfund site). Figure 6-1 illustrates the underlying
differences between raster and vector approaches for organizing aspects of the
real world into the digitized features contained in GIS data layers. Table 6-1
compares the advantages and disadvantages and recommends uses of raster-
and vector-based GIS programs.
Although there was formerly a major divergence between GIS systems designed
to handle raster as opposed to vector data layers, most GIS packages now will
either contain procedures for handling both data types or provide transformation
programs that can convert one format to the other. While raster-based systems
have advantages when dealing with information such as land cover or soil types
over large geographic areas, vector approaches have become increasingly
popular for most routine GIS analysis applications.
To convert real-world information into GIS data layers, important objects and
features must be located precisely so that different data layers will overlay
correctly. Geographic information contains either an explicit geographic
reference, such as a latitude and longitude or national grid coordinate, or an
implicit reference, such as an address, postal code, census tract name, or road
name. An automated process called geocoding is used to create explicit
geographic references from implicit references (descriptions such as addresses).
These geographic references allow you to locate features, such as a Superfund
site, and events, such as the location of a major chemical spill, on the earth's
surface for analysis. In the vector model, information about points, lines, and
6-2
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6. MAPPING TOOLS
Raster
Vector
Real
World
Figure 6-1. GIS data layers may use raster or
vector representation techniques.
polygons is encoded and stored as a collection of x,y coordinates. The location
of a point feature, such as a point source discharge, can be described by a single
x,y coordinate. Linear features, such as roads and rivers, can be stored as a
collection of point coordinates. Polygonal features, such as watershed catchments
or the boundaries of political units, such as towns, can be stored as a closed loop
of coordinates.
The geocoding process can be the most time-consuming and resource-intensive
step in a GIS analysis and mapping process. Data layers involving point or poly-
gon features can be especially difficult to digitize to high degrees of precision. On
the other hand, point coverages are often much easier to create. For point
coverage, the main requirements are an accurate set of latitude and longitude
coordinates or locational information from global positioning satellite (GPS) tools.
Point data layers (or coverages) can also be created using existing line or polygon
coverages as base maps, from which the point locations can be supplied using
software tools in a GIS.
A sensible strategy in conducting special risk analysis or risk management
projects with GIS is to identify what data layers are already available and keep the
coverages that must be created from scratch to a minimum. The new coverages,
in many cases point coverages, would be based on site-specific information
based on special surveys or data collections. For existing coverages or
georeferenced data files, facilities accessible through the Internet
6-3
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6. MAPPING TOOLS
Table 6-1. Comparison of Raster- Versus Vector-Based GIS Programs
Raster Method Vector Method
Advantages
Disadvantages
Recommended
Uses
Simple data structure
Overlay and combination of
mapped data with remotely
sensed data is easy
Various kinds of spatial analyses
are easy
Simulation is easy because each
spatial unit has the same size
and shape
Technology is inexpensive and
is being actively developed
Volumes of graphic data
Use of large cells to reduce data
can lose important data, so
frequently cannot simplify
information
Raster map graphics are more
crude than vector maps drawn
with fine lines
Network linkages are difficult to
establish
Projection transformations are
time consuming unless special
algorithms or hardware is used
Quick and inexpensive overlay,
map combination and spatial
analyses
Simulation and modeling when
working with surfaces is
necessary
Good representation of
phenomena (such as county
and towns, or soil structure
hierarchies)
Compact data structure
Topology can be described
completely with network
linkages
Retrieval, updating, and
generalization of graphics and
attributes are possible
Complex data structures
Combination of several vector
maps through overlay creates
difficulties
Simulation is difficult because
each unit has a different
topological form
Display can be expensive,
particularly for high quality,
color, and cross-hatching
Technology is expensive,
especially for more
sophisticated software and
hardware
Spatial analyses and filtering
within areas are impossible
Data-archiving phenomena
(e.g., soil areas, land use units)
Network analyses (e.g.,
telephone networks or
transportation networks)
Compact digital terrain models
Source: Burrough (1991).
and the World Wide Web (WWW or Web) are making it easier to locate and
obtain (often for free) a variety of useful data products. Major impetus for using
the Internet to exchange GIS data has come from the federal initiative known as
the National Spatial Data Infrastructure (NSDI). EPA has strongly supported this
effort and, in partnership with other federal and state agencies, now offers a broad
spectrum of valuable data products through its Web pages.
6-4
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6. MAPPING TOOLS
6.4 INTERNET SOURCES OF EXISTING DATA FILES AND CIS COVERAGES
A consortium of major governmental agencies cooperates through the Federal
Geographic Data Committee (FGDC) to encourage the widest possible use of
good quality spatial data products. The main mechanism for sharing these
information products is through a series of special Internet facilities maintained by
individual federal or state agencies, university research groups, and NSDI. The
NSDI is conceived to be an umbrella of policies, standards, and procedures under
which organizations and technologies interact to foster more efficient use,
management, and production of geospatial data. The Clinton Administration has
asked the FGDC to provide the federal leadership for evolving the NSDI in
cooperation with state and local governments and the private sector.
The Internet provides a number of interactive software tools to share information,
but the most popular tools center on the use of Web browsers that are available
for computers of all types ranging from sophisticated workstations to personal
computers. A growing number of private citizens use Web browsers at their
homes. The URL providing general information for the entire NSDI is:
.
This central hub for the NSDI provides Web links to a number of other major
"nodes" in the NSDI system. Federal agencies such as the Census Bureau, the
United States Geological Survey (USGS), the USDA, and EPA have their own
NSDI Web pages with links to more specialized data items. EPA's link to the NSDI
is at .
EPA has also established a number of Web pages to help provide background
information or help access actual data products dealing with particular databases
or EPA programs. Examples include a facility called "Surf Your Watershed," which
acts as a gateway to information organized according to standard watershed
catchments called Hydrologic Cataloging Units defined by the USGS, and an
Internet data warehouse system called ENVIROFACTS that allows the retrieval
of information dealing with permitted facilities (e.g., Permit Compliance Systems
[PCS] for point source discharges to receiving waters), Superfund, or
Comprehensive Environmental Response, Compensation, and Liability Act List
of Sites (CERCLIS), and information from databases such as the TRI.
With the EPA Web facilities, data files or GIS coverages may be downloaded that
could then be incorporated into risk assessment and management projects; the
end user would then need access to a GIS to perform spatial analyses and
produce the final GIS maps. EPA is also setting up Web facilities at which the
user can provide inputs on the type of analysis to perform and then retrieve maps
directly from the Internet link. An example is given in Figure 6-2 of a Web tool
called BASININFO that can produce displays of the major types of permitted
facilities within a USGS Cataloging Unit. Several WEB-based data retrieval and
mapping tools are now part of EPA's Maps On Demand systems, which can be
accessed at the following address:
.
6-5
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6. MAPPING TOOLS
B VSIMIVFO Map Dalii anil Output Selection
3EF9V
LEGEND
Population Density Par 85 Mi
~' Under 3.MO
'K'1 QW JO.**
t 8 3 i S i f ! 9
Figure 6-2. Examples of GIS displays from EPA's
BASININFO Maps-on-Demand facility.
EPA's SU RF YOU R WATERSH ED facility provides an on-line set of maps derived
from the Office of Science and Technology's National Listing of Fish and Wildlife
Advisories (NLFWA) Database. Figure 6-3 shows a display depicting the locations
of active advisories for the State of North Carolina. GIS maps showing the
location offish advisories in any of the 50 states, U.S. territories, and the District
of Columbia can be viewed on this system.
6.5 DATA NEEDED FOR MAPPING
The information needed for a given map depends largely on the objective of the
map itself. The following major categories of information are useful for mapping:
• Chemical contaminant type and concentration
• Consumer population
• Risk level.
Additional refinements may be desirable, including the relationship of chemical
contaminants to various point or nonpoint sources, demographic characteristics
of the consumer population, consumption patterns of population groups, and
types and levels of human health risks. At a minimum, contaminant mapping is
usually possible because sampling and analysis programs are basic to all fish
advisory programs and generate the necessary data to map the locations where
various contaminants are detected as well as the fish species and size (age class)
in which the contaminant occurs. Individual maps for each contaminant may be
generated, or maps of several contaminants can be displayed together if there is
6-6
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6. MAPPING TOOLS
Richmond T i$ /i
'
r- Raleigh
- Columbia
Figure 6-3. Map showing active fish and wildlife advisories for a state.
sufficient refinement in the system. Contaminant concentration can be indicated
using different colors; through graphic patterning such as cross-hatching, lines,
and dots; or through the use of different symbols (open, semiclosed, or closed
circles or squares).
6.6 MAPPING PROGRAMS
Computerized mapping programs are useful aids; however, mapping programs
take some time to learn and require data collection and organization prior to data
entry. State and local agencies interested in digital mapping should consider the
following:
• Availability of the data needed for each map
• Quality of the data to be used
• Amount of time and money available
• Type of program used to generate maps
• Purpose of each map or map series for developing consumption advisories.
It is important to evaluate the goals of the mapping effort and the resources
available for the activity. Using a program that does more than is needed can
result in unnecessary expenditures for staff training and developing maps for
analysis. Data storage capacity is also an important consideration and may be a
factor in choosing a mapping approach.
6-7
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6. MAPPING TOOLS
Many federal, regional, state, and tribal agencies already have some divisions that
are using GIS programs for other purposes. It is cost- and time-effective to consult
with staff already using this resource. Several mapping programs are available
that are relatively uncomplicated and inexpensive. These programs are often
called desktop mapping or desktop GIS packages. One example of a commercial
desktop GIS package is ESRI's ArcView, which can be set up on a personal
computer. Generally, PC-based programs can be used to digitize field map data
onto a computer, but these programs often have limited capacity to accommodate
large data sets. Although more sophisticated programs that usually require high-
performance workstations as their computer platforms offer greater flexibility in
data input and manipulation, they are often an expensive option and require more
expertise to set up and operate. Most GIS programs can generate large volumes
of data that need to be stored, so consider computer space in advance.
One cost-effective and sophisticated program, run as a nonprofit venture, has
been used extensively by international nongovernmental organizations (NGOs)
and intergovernmental organizations with great success. IDRISI (whose name is
taken from a medieval Arabic geographer who lived in what is now Morocco) is
available from the Geography Department of Clark University in Massachusetts.
It consists of inexpensive software that can use and manipulate data easily and
also be programmed to assist in selecting outlining criteria for management
analyses. The University offers training workshops and other assistance for new
users (including Applications in Forestry, Coastal Zone Research and
Management, and Decision Making), which may be useful for fish advisory
program staff. The IDRISI program is a raster-based system, so the analyses
conducted by the program are performed rapidly, effectively, and relatively
inexpensively. This particular program is sophisticated enough to accommodate
some of the more complicated analyses that are normally difficult to perform
without a vector-based program.
Mapping information for the development and management of fish advisories is
a relatively new undertaking for most agencies. EPA welcomes ideas and
recommendations on this topic. Examples of maps or mapping methods that are
widely applicable are especially welcome.
6.7 NATIONAL LISTING OF FISH AND WILDLIFE ADVISORIES (NLFWA) DATABASE
Mapping information for the development and management of fish advisories is
a relatively new undertaking that provides precise information to fish-consumers
or those waterbodies where chemical contamination in fish may be of public
health concern for most agencies.
The EPA Office of Science and Technology within the Office of Water has
developed a new Internet Web-based platform for the NLFWA database. State,
regional, and local governmental staff as well as members of the general public
can now search this database to obtain narrative information on fish consumption
advisories and bans. In addition, users can also electronically retrieve and print
state, regional, and national maps showing the geographic location and extent of
6-8
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6. MAPPING TOOLS
the fish advisories in each of the states, District of Columbia, and the four U.S.
territories.
Information on the geographic extent of the advisories and bans is provided to
EPA by the states either as narrative information (e.g. advisory includes all waters
of the Black River from its source to the stateline), as latitude and longitude
coordinates, or hand-marked on USGS maps which can be digitized into the
Geographic Information System (GIS). Other tools used to find the location of
waterbodies under fish advisory include: searchable CD-ROMs of geographic
information such as TopoUSA, digital tables of geographic sites provided by
USGS, and other county maps or information (such as the county where an
advisory occurs, or length of advisory) in state reports and memorandums. Tribal
authorities also provided computer-generated maps in reports and an electronic
GIS file of U.S. waterways include the names of waterbodies that can be
searched and matched to information provided by the states.
This GIS mapping information related to fish consumption advisories and bans is
available on the Internet at:
http://www.epa.gov/ost/fish
State, regional, and local agency staff may obtain additional information on the
new Internet WEB-based database EPA now has available by contacting:
U.S. Environmental Protection Agency
Office of Science and Technology
National Fish and Wildlife Contamination Program - 4305
1200 Pennsylvania Avenue, NW
Washington, DC 20460
PHONE: 202-260-7301
FAX: 202-260-9830
E-Mail: bigler.ieff@epa.gov
6-9
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7. LITERATURE CITED
SECTION 7
LITERATURE CITED1
Abernathy, C.O., R. Cantilli, J.T. Du, and O.A. Levander. 1993. Essentiality
versus toxicity: Some considerations in the risk assessment of essential trace
elements. In: Vol 8, Hazard Assessment of Chemicals, J. Saxena (ed.),
Taylor and Francis, Washington, DC.
Abernathy, C.O., and W.C. Roberts. 1994. Risk assessment in the Environmental
Protection Agency. J. Haz. Mat. 39(2): 135-142.
Anderson, H.A., and J.F. Amrhein. 1993. Protocol for a Uniform Great Lakes
Sport Fish Consumption Advisory. Prepared for the Great Lakes Advisory
Task Force. May.
ATSDR (Agency for Toxic Substances and Disease Registry). 1990. Toxicological
Profile for Endrin. U.S. Department of Health and Human Services, Public
Health Service, Atlanta, GA.
ATSDR (Agency for Toxic Substances and Disease Registry). 1991. Draft
Toxicological Profile for Dieldrin. U.S. Department of Health and Human
Services, Public Health Service, Atlanta, GA.
ATSDR (Agency for Toxic Substances and Disease Registry). 1992.
Toxicological Profile for Tin and Tin Compounds. U.S. Department of Health
and Human Services, Public Health Service, Atlanta, GA.
ATSDR (Agency for Toxic Substances and Disease Registry). 1993a.
Toxicological Profile for Endosulfan. U.S. Department of Health and Human
Services, Public Health Service, Atlanta, GA.
ATSDR (Agency for Toxic Substances and Disease Registry). 1993b.
Toxicological Profile for Heptachlor Epoxide. U.S. Department of Health and
Human Services, Public Health Service, Atlanta, GA.
1 Article titles were not usually available for citations obtained from HSDB; consequently, page numbers
were included for those citations (only).
7-1
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7. LITERATURE CITED
ATSDR (Agency for Toxic Substances and Disease Registry). 1994a. Draft
Toxicological Profile for Chlordane. U.S. Department of Health and Human
Services, Public Health Service, Atlanta, GA.
ATSDR (Agency for Toxic Substances and Disease Registry). 1994b. Draft
Toxicological Profile for ODD, DDT, DDE. U.S. Department of Health and
Human Services, Public Health Service, Atlanta, GA.
ATSDR (Agency for Toxic Substances and Disease Registry). 1994c. Draft
Toxicological Profile for alpha, beta, gamma, and delta
Hexachlorocyclohexane. U.S. Department of Health and Human Services,
Public Health Service, Atlanta, GA.
ATSDR (Agency for Toxic Substances and Disease Registry). 1995a.
Toxicological Profile for Mirex and Chlordecone. U.S. Department of Health
and Human Services, Public Health Service, Atlanta, GA.
ATSDR (Agency for Toxic Substances and Disease Registry). 1995b.
Toxicological Profile for Polycyclic Aromatic Hydrocarbons (PAHs). U.S.
Department of Health and Human Services, Public Health Service, Atlanta,
GA.
ATSDR (Agency for Toxic Substances and Disease Registry). 1996a.
Toxicological Profile for Selenium. U.S. Department of Health and Human
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ATSDR (Agency for Toxic Substances and Disease Registry). 1996b.
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7-10
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7. LITERATURE CITED
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7-11
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7. LITERATURE CITED
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7-12
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7. LITERATURE CITED
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7-13
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7. LITERATURE CITED
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Research and Development, Washington, DC. August.
U.S. EPA (Environmental Protection Agency). 1997c. Reference Dose
Tracking Report. Office of Pesticide Programs, Health Effects Division,
Washington, DC.
U.S. EPA (Environmental Protection Agency). 1997d. Mercury Study Report to
Congress. EPA-452R-96-001a-h. Office of Air Quality Planning and
Standards, Research Triangle Park, NC, and Office of Research and
Development, Cincinnati, OH.
U.S. EPA (Environmental Protection Agency). 1997e. Toxicological Review of
Chlordane (Technical). Office of Pesticide Programs, Washington, DC.
December.
U.S. EPA (Environmental Protection Agency). 1997f. Exposure Factors
Handbook, Volume II. Food Ingestion Factors. EPA/600/P-95/002Fb.
Office of Research and Development, Washington, DC. August.
U.S. EPA (Environmental Protection Agency). 1997g. Toxicological Review of
Tributyltin Oxide (CAS No. 56-35-9) in Support of Summary Information on
the Integrated Risk Information System (IRIS). Washington, DC. July.
7-14
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7. LITERATURE CITED
U.S. EPA (Environmental Protection Agency). 1997h. Terbufos-FQPA
Requirement-Report of the Hazardous Identification Review. Office of
Pesticide Programs. Washington, DC.
U.S. EPA (Environmental Protection Agency). 1998a. Reregistration Eligibility
Decision (RED). Dicofol. Office of Pesticide Programs and Toxic
Substances, Washington, DC.
U.S. EPA (Environmental Protection Agency). 1998b. Memorandum dated
April 1, 1998, Diazinon: Report of the Hazard Identification Assessment
Review Committee. HED DOC NO. 012558. Office of Pesticide Programs,
Health Effects Division.
U.S. EPA (Environmental Protection Agency). 1998c. Memo dated September
24, 1998. Reused Oxyfluorfen (Goal) Quantitative Risk Assessment (Q1*)
Based on CD-1 Male Mouse Dietary Study with 3/4's Interspecies Scaling
Factor. HED DOC NO. 012879. Office of Pesticide Programs, Health
Effects Division.
U.S. EPA (Environmental Protection Agency). 1999a. Guidance for Conducting
Health Risk Assessment of Chemical Mixtures. NCEA-C-0148. Risk
Assessment Forum Technical Panel, Office of Research and Development,
Office of Pesticide Programs, Office of Pollution Prevention and Toxics,
and Office of Water, Washington, DC. April.
U.S. EPA (Environmental Protection Agency). 1999b. Recognition and
Management of Pesticide Poisonings, 5th Ed. U.S. Government Printing
Office, Washington, DC.
U.S. EPA (Environmental Protection Agency). 1999c. List of Chemicals
Evaluated for Carcinogenic Potential. Office of Pesticide Programs.
Health Effects Division, Washington, DC.
U.S. EPA (Environmental Protection Agency). 1999d. Human Health Risk
Assessment - Ethion. Office of Pesticide Programs, Health Effects
Division, Washington, DC. July 14.
U.S. EPA (Environmental Protection Agency). 2000a. Guidance for Assessing
Chemical Contaminant Data for Use in Fish Advisories. Volume 1: Fish
Sampling and Analysis, Third Edition. Office of Science and Technology,
Office of Water, Washington, DC.
U.S. EPA (Environmental Protection Agency). 2000b. Revised Human Health
Risk Assessment for Chlorpyrifos. Office of Pesticide Programs,
Washington, DC.
7-15
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7. LITERATURE CITED
U.S. FDA (Food and Drug Administration). 1993. Guidance Document for
Cadmium in Shellfish. Center for Food Safety and Applied Nutrition,
Rockville, MD.
U.S. FDA (Food and Drug Administration). 1998. Action Levels for Poisonous
or Deleterious Substances in Human Food and Animal Feed. Industry
Activities Staff Booklet. Washington, DC.
Van den Berg, M., L. Birnbaum, A.T.C. Bosveld et al. 1998. Toxic equivalency
factors (TEFs) for PCBs, PCDDs, PCDFs for human and wildlife.
Environmental Health Perspective 106(12):775-792.
Velazzquez, Susan. 1994. Personal communication. U.S. EPA. Environmental
Criteria and Assessment Office, Cincinnati, OH.
Voiland Jr., M.P., K.L. Gall, D.J. Lisk, and D.B. MacNeill. 1991. Effectiveness
of recommended fat trimming procedures on the reduction of PCB and
mirex levels in brown trout (salmo trutta) from Lake Ontario. J Great Lakes
ftes17(4):454-460.
West, P.C., M.J. Fly, R. Marans, and F. Larkin. 1989. Michigan Sports Anglers
Fish Consumption Survey, Supplement I, Non-Response Bias and
Consumption Suppression Effect Adjustments. School of Natural
Resources, University of Michigan, Ann Arbor. Natural Resource Sociology
Research Lab, Technical Report No. 2.
WHO. 1990. Environmental Health Criteria 101: Methylmercury. Geneva,
Switzerland: WHO.
WHO. 1999. Concise International Chemical Assessment Document No 14:
Tributyltin Oxide. Geneva, Switzerland: WHO.
Wulf, H.C., N. Kromann, N. Kousgaard, etal. 1986. Sister chromataid
exchange (SCE) in Greenlandic Eskimos: Dose-response relationship
between SCE and seal diet, smoking, and blood cadmium and mercurcy
concentrations. Sci Total Environ. 48:81-94.
7-16
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APPENDIX A
REVIEWERS OF FIRST EDITION OF GUIDANCE DOCUMENT
-------
APPENDIX A
APPENDIX A
Reviewers of First Edition of Guidance Document
The following individuals, representing EPA Headquarters, EPA Regions, state and
federal agencies, and Native American groups provided technical information, reviews, and
recommendations throughout the preparation of the first edition. Participation in the review
process does not imply concurrence by these individuals with all concepts and methods
described in this document.
EPA Headquarters
Jeffrey Bigler
Charles Abernathy
Tom Armitage
Kenneth Bailey
Denis Borum
Robert Cantilli
James Cogliano
Joyce Donohue
Julie Du
Rick Hoffmann
Skip Houseknecht
Frank Gostomski
Amal Mahfouz
Bruce Mintz
Edward Ohanian
Rita Schoeny
Betsy Southerland
Margaret Stasikowski
Yogi Patel
William Farland
Gregory Kew
Carole Kimmel
Gary Kimmel
Jackie Moya
Lorenz Rhomberg
Reto Engler
George Ghali
Michael Metzger
Esther Rinde
Steve Shaible
Richard Whiting
EPA/Office of Water (Workgroup Chairman)
EPA/Office of Water
EPA/Office of Water
EPA/Office of Water
EPA/Office of Water
EPA/Office of Water
EPA/Office of Water
EPA/Office of Water
EPA/Office of Water
EPA/Office of Water
EPA/Office of Water
EPA/Office of Water
EPA/Office of Water
EPA/Office of Water
EPA/Office of Water
EPA/Office of Water
EPA/Office of Water
EPA/Office of Water
EPA/Office of Water
EPA/Office of Research and Development
EPA/Office of Research and Development
EPA/Office of Research and Development
EPA/Office of Research and Development
EPA/Office of Research and Development
EPA/Office of Research and Development
EPA/Office of Pesticide Programs
EPA/Office of Pesticide Programs
EPA/Office of Pesticide Programs
EPA/Office of Pesticide Programs
EPA/Office of Pesticide Programs
EPA/Office of Pesticide Programs
A-3
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APPENDIX A
Other EPA Office Staff
Eletha Brady-Roberts
John Cicmanec
Michael Dourson
Susan Velazquez
Chon Shoaf
Jerry Stober
Charles Kanetsky
Milton Clark
Philip Crocker
EPA/Office of Research and Development, Cincinnati, OH
EPA/Office of Research and Development, Cincinnati, OH
EPA/Office of Research and Development, Cincinnati, OH
EAP/Office of Research and Development, Cincinnati, OH
EPA/Environmental Criteria and Assessment Office, RTP.NC
EPA/Environmental Research Laboratory, Athens, GA
EPA/Region 3
EPA/Region 5
EPA/Region 7
Other Federal Agency Staff
Michael Bolger
Gregory Crame
Gunnar Lauenstein
Thomas Siewicki
Janice Cox
U.S. Food and Drug Administration
U.S. Food and Drug Administration
National Oceanic and Atmospheric Administration
National Oceanic and Atmospheric Administration
Tennessee Valley Authority
State Agency Staff
Anna Fan
Gerald Pollock
Richard Greene
Joseph Sekerke
Tom Long
Dierdre Murphy
Jack Schwartz
John Hesse
Pamela Shubat
Gale Carlson
Alan Stern
Robert Tucker
California
California
Delaware
Florida
Illinois
Maryland
Massachusetts
Michigan
Minnesota
Missouri
New Jersey
New Jersey
Tony Forti
Luanne Williams
Martin Schock
Kandiah Sivarajah
Robert Marino
Kim Blindauer
Alan Anthony
Peter Sherertz
Ram Tripathi
Denise Laflamme
Jim Amrhein
Henry Anderson
New York
North Carolina
North Dakota
Pennsylvania
South Carolina
Utah
Virginia
Virginia
Virginia
Washington
Wisconsin
Wisconsin
Native American Tribes
Neil Kmaicik
Ann Watanabe
John Banks
Clemon Fay
Great Lakes Indian Fish and Wildlife Commission
Columbia River Inter-Tribal Fish Commission
Penobscot Nation
Penobscot Nation
A-4
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APPENDIX B
POPULATION EXPOSURE ASSESSMENT-
CONSUMPTION PATTERNS AND SURVEYS
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APPENDIX B
APPENDIX B
POPULATION EXPOSURE ASSESSMENT-CONSUMPTION PATTERNS
AND SURVEYS
Selecting appropriate population exposure data is critical in both risk estimation
and in fish advisory program planning. Whenever possible, state agencies are
encouraged to conduct local surveys to obtain information on consumption
patterns. The time and resources required to conduct onsite surveys, however,
can be prohibitive. If only limited local data are available, that information may be
used and supplemented with the best available data from other sources. If local
or regional data are not available and surveying is not feasible, other sources may
be used to characterize the consumption patterns of a population.
B.1 HIERARCHY OF FISH CONSUMPTION INFORMATION
Table B-1 lists a hierarchy of information sources on fish consumption that may
be considered in obtaining data for developing fish advisories. Care should be
taken when selecting a matched population and consumption data set to use as
"representative" of the target population. Matches should be made based on
similar consumption patterns, rather than on generalizations about ethnic behavior
or other attributes.
Matching groups with high consumption rates to previously studied groups having
similar characteristics is particularly important. These groups with high
consumption rates are often those of greatest concern due to their higher potential
risks. They are at greater risk than the general population if their consumption is
underestimated and may also be more severely jeopardized by losing their fish
food sources than the general population if their consumption rates are
overestimated.
Many studies are not appropriate for use in exposure assessment. Surveys may
be based on only those fishers who apply for licenses through state agencies; this
often underestimates consumption rates in some subpopulations. In some areas,
the results may reflect a combination of commercially caught fish as well as
subsistence- or sport-caught fish and may therefore provide an incomplete picture
offish consumption patterns in a particular region. Often, qualitative or anecdotal
information is available to corroborate or challenge the results of older data; this
can help to assess the need for additional data collection. For example, a survey
may have been conducted in a state with a large urban Asian-American
population, commonly known to eat large quantities of fish, yet only a small
B-3
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APPENDIX B
Table B-1. Hierarchy of Data Sources3
1. Local fish consumption survey (creel surveys)
2. Local fish consumption survey with limited scope
(e.g., acquired by fish licenses only)
3. Regional or state survey data from other areas having matching characteristics"
• Behavioral Risk Surveillance Survey (BRSS)
• Anecdotal information
4. National fish or food consumption data taking into consideration demographic data
• National Survey of Fishing, Hunting, and Wildlife Associated Recreation (U.S. Fish
and Wildlife Service, 1993)
• U.S. Department of Agriculture Continuing Survey of Food Intake by Individuals
(CSFII) studies
• Other national surveys that estimate fish consumption patterns
• Census data
3 This hierarchy is generally applicable; however, the utility of any data source is dependent on the
match between the population studied in the data source and that being considered by the risk
managers. For example, when a better match is available through national or regional fish con-
sumption data than can be found through limited local fish surveys, then the national, regional, orstate
data are preferable. Special care should be taken that data for highly exposed subpopulations are
obtained from sources that considered populations with equally high exposures.
b Secondary data sources can be used most effectively in conjunction with qualitative data and
anecdotal information (e.g., informal discussions with community groups, clerks, and other qualitative
studies).
number of the survey respondents were Asian-American. If the survey was
conducted by fishing license registration, it is likely that a large portion of the
exposed population was unintentionally excluded from the survey and thus was
not adequately represented in the consumption estimates.
B.1.1 Local Fish Consumption Data
B.1.1.1 Creel Surveys—
Another source of information concerning fishing habits (applicable indirectly to
consumption estimates) is obtained through the creel surveys. Most state
B-4
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APPENDIX B
agencies involved with fish and wildlife management perform creel surveys or
censuses. These surveys consist of clerks interviewing fishers onsite and
recording the size and species of fish they take home (and presumably eat).
These surveys are performed to calculate fishing pressures and evaluate stocking
programs for state lakes and streams. These surveys generally contain little
demographic information beyond the fisher's home county, though they may be
modified to ask additional questions about demographics and fish consumption.
Creel surveys are subject to reporting biases, which may include a reluctance of
fishers to report a poor catch or a catch that exceeds allowable limits (see a
discussion of data collection problems below). The clerks themselves know a
great deal of anecdotal information about fishers because of their direct contact
with these individuals. Clerks, area fisheries managers, and conservation officers
are excellent sources of information on fisher demographics and should be
contacted during research into most fisher populations (Shubat, 1993). Like
surveys taken only from licensed fishers, however, this qualitative information may
be restricted to certain fishers and fishing locations.
B.1.1.2 Fishing License Surveys-
Fishing license tracking may be a good source for obtaining demographic
information for target populations. Fishing licenses include information on the
name, age, and address of fishers, location where the license was sold, and the
approximate length of the fishing trip (e.g., 4-day, seasonal). Although the
information on the license is limited, some researchers have used the addresses
on licenses to send out more detailed surveys. Several fish advisory programs,
including those in Minnesota and Canada, insert detailed demographic and
consumption surveys in their informational booklets, which fishers may fill out and
return in exchange for receiving the following year's materials. These surveys by
definition, however, reach only a portion of respondents already aware of the fish
programs (Shubat, 1993). They also do not reach fishers who do not purchase
licenses for economic or other reasons. In addition, Native American groups who
are often legally entitled to fish on tribal waterbodies without licenses will not be
accessed by this method.
B.1.2 Regional or State Consumption Data
B.1.2.1 Anecdotal Information-
Anecdotal information is vital in directing the search for data on fish consumption
patterns. For example, anecdotal information suggests that urban and rural fishers
often sell their products "informally" (i.e., without commercial licenses) in
geographic areas near where they fish and have customers with "standing orders"
for regular fish delivery. This practice has been observed in Missouri, Mississippi,
Alaska, and in the Chicago and Milwaukee metropolitan areas and is common to
both rural and urban areas (Carlson, 1994). Health officials have raised concerns
that "customers," who tend to be from minority or low-income populations, may
B-5
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APPENDIX B
be exposed to contaminant concentrations over a long period of time. These
groups, while not composed entirely of fishers, may have exposure levels as high
as those for subsistence fishers (Carlson, 1994). Another exposed group that may
not be well-characterized in some surveys is made up of fishers' family members,
including extended families to whom fish is supplied.
Under these circumstances of unlicensed distribution it is likely that
• Those consuming the fish are unaware of the fish advisories, even if the
actual fisher is aware
• Contacting the fisher is often difficult and the fisher, once reached, may be
very reluctant to provide data on fish catch rates for fear of prosecution.
To obtain an estimate of consumption occurring via these routes, information can
be acquired through informal discussions with local community groups in areas
of potential exposure.
B.1.2.2 Behavioral Risk Surveillance Surveys-
Most states already participate in random telephone surveys under the Behavioral
Risk Surveillance System (BRSS). The BRSS surveys are often the only random,
state-level survey information readily available to states. They are funded by the
Agency for Toxic Substances and Disease Registry (ATSDR), a department within
the Center for Disease Control and Prevention (CDC). Some states have already
used federal grant money to add questions on fisher demographics and
consumption to the BRSS surveys (Shubat, 1993).
B.1.3 National Consumption Data
B.1.3.1 National Survey of Fishing, Hunting and Wildlife—
The U.S. Fish and Wildlife Service (FWS) conducts a survey every 5 years that
includes data on sport fishing. The most recent survey is entitled 1991 National
Survey of Fishing, Hunting and Wildlife Associated Recreation (\J.S. FWS, 1993)
and is available from the FWS. This survey provides information by state on
fishers, broken down by age, sex, race/ethnic group, and state of residence. The
FWS data can be used in combination with local data on the size of the fishing
population overall to estimate the numbers of exposed individuals with relevant
exposure characteristics. For example, using the FWS data, one could estimate
the percentage of fishers in the state in a certain age group and apply this
percentage to local fishing population data (from fishing licenses, for example) to
estimate the number of local fishers in that age group.
B.1.3.2 U.S. Department of Agriculture (USDA) CSCFII Study-
B-6
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APPENDIX B
The Continuing Survey of Food Intake by Individuals (CSFII) is a national food
consumption survey conducted annually by the USDA. It consists of multistage,
stratified-area probability samples from all states except Alaska and Hawaii. In the
CSFIIs, dietary intake data collection is distributed over a year-long period. Survey
participants provide 3 consecutive days of data. On the first day of the survey,
participants provide information to an in-home interviewer. On the second and
third days, data are taken from self-administered dietary records. Meals
consumed both at home and away from home are recorded (U.S. EPA, 1998b).
B.2 FISH CONSUMPTION SURVEY METHODS
If time and money permit, researchers are encouraged to conduct their own
surveys to characterize fisher populations. EPA's guidance manual, Guidance for
Conducting Fish and Wildlife Consumption Surveys (U.S. EPA, 1998a) may be
useful in planning demographic surveys. Researchers also may consider
coordinating survey efforts with other existing programs. For example, many state
agencies conduct educational outreach programs to provide information or explain
new regulations to fishers. Health agencies and natural resource offices can
combine efforts to target subpopulations not yet reached through other
mechanisms.
B.2.1 Key Considerations
Table B-2 lists key considerations in conducting effective fish consumption
surveys. Although surveying of a specific population can provide the most
accurate exposure information about it, care must be taken in conducting the
survey. The credibility of the survey results must be ensured through careful
survey preparation, sample selection, and administration.
Population selection is one of the most significant components of an exposure
assessment. A tiered approach is a logical recommendation for selecting popula-
tions of concern. First, examine the areas surrounding waterbodies that have
been identified as contaminated or supporting potentially contaminated fish (e.g.,
anadromous fish arriving from contaminated estuaries).
Following this range identification, collect as much anecdotal information as
possible from local populations surrounding these waterbodies. Qualitative data
will indicate what communities are supported by the waterbodies, whether people
are traveling long distances to fish in the waters, and other useful information to
help direct further steps of the consumption evaluations. At this point, review the
following information to determine whether a further investigation should be
carried out:
• Anecdotal information suggesting high consumption rates
• Fish consumption patterns indicating potentially high exposure
B-7
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APPENDIX B
Table B-2. Key Considerations for Effective
Fish Consumption Surveys
Population Selection What population is to be surveyed?
Based on what criteria (e.g., jurisdictional region, region with
known fish contamination)?
Population Access How will the identified population be reached?
Will separate methods be used for distinct subpopulations
(e.g., fish licensing for sport fishers, community groups for
urban subsistence fishers)?
Consumption Rates What method will be used to estimate consumption rates
(e.g., recall, recordkeeping, catch rate)?
What assumptions are made in these estimations (e.g., meal
size, household size)?
Consumption Patterns How are variations in consumption patterns accommodated
(e.g., preparation methods, type offish eaten, parts offish
consumed)?
Duration of Study Have consumption rates been estimated for each different
season or generalized?
Have large fish catches that have been frozen or preserved
for nonfishing seasons been addressed?
• Subpopulations known to have high consumption rates living in the region or
identified as fishing in the waters of concern, whether or not any anecdotal
evidence exists to support high consumption or exposure rates.
Once the target population is selected, some method must be chosen to survey
these individuals. As mentioned earlier, using fishing licenses as a survey tool
may miss a large portion of the fishing population. It may be most useful to enlist
the help of local agencies or community groups to help access some of the
subpopulations at high risk, such as urban low-income populations or individuals
of a particular ethnicity. Both identifying populations and collecting data may rely
heavily on qualitative or anecdotal evidence on fishers to evaluate exposures of
highly exposed populations. Consumption patterns affecting the overall
consumption rate and toxicity must be discerned as well, including:
• Species of fish consumed
• Portions of fish that are consumed (fillet only or whole body)
• Preparation and cooking methods.
B-8
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APPENDIX B
A determination must be made as to whether fish is a major source of protein in
the diet of the subpopulation of concern. If advisories are developed based on the
survey results, this information can provide some clue about the impact of fishing
restrictions as one risk management option.
Several methods can be used to estimate a population's consumption rate. Actual
recordkeeping for some period of time is the most accurate method, although a
long-term commitment is needed from the respondents. Memory-recall is another
method used to estimate consumption rates. This method can take the form of
either "how many meals offish (or what amount offish) have you (and household
members) eaten in this past week'?" or "how many meals offish (or what amount
offish) do you (and household members) eat each week in general?" While the
length of recall can vary, long-term recall introduces uncertainties and
inaccuracies. Individuals knowing the objective of the survey may be biased in
their memory recall as well.
Meal size is another feature of determining consumption patterns. Many fish
advisories are developed based on assumptions regarding meal size or specific
consumption limits for a specific meal size. If information is not collected on meal
size, risk managers may wish to use the average meal size assumption
recommended by EPA of 227 g (8 oz) of fillet per 70 kg consumer body weight for
adults. This value has been cited as appropriate in many documents on fish
consumption (Anderson and Amrhein, 1993; Dourson and Clark, 1990; Minnesota
Department of Health, 1992; Missouri Department of Health, 1992; U.S. EPA,
1988,1995). This 8-oz fish meal weight may be considered an average meal size.
For those populations who consume fish whole, or who consume nonfilleted
portions of the fish, meal sizes should be obtained from qualitative data or direct
surveys. Readers are urged to collect information on meal size specific to their
areas and populations of concern, especially if very large meals are known to be
consumed during fishing trips, festivals, or under other circumstances. Information
regarding maximum meal size may also be valuable in determining whether risks
are likely to arise from large short-term exposures (bolus doses).
B.2.2 Data Collection Problems
Conducting surveys to assess the consumption of noncommercially caught fish
can be particularly challenging. Numerous individuals involved with fish consump-
tion surveys have raised issues not mentioned in prior guidance documents. Their
most notable concern was that of assessing the consumption rates of urban
fishers or minority groups that were not registered for fishing licenses. In addition,
surveys were often returned with consumption rates that were inconsistent with
observed habits and the available qualitative data.
Surveys conducted using traditional methods can exclude major portions of the
fish-consuming population. Several localities have attempted to conduct surveys
to more accurately reflect the true consumption patterns existing within each
B-9
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APPENDIX B
subpopulation. However, they found that, in some cases, unregistered fish
consumers were answering survey questions inaccurately for any number of
reasons, including the following:
• Fishers associated the state or local agency conducting the survey with
enforcement and provided responses they thought the surveyors wanted to
hear.
• Individuals who run illegal fish markets and are afraid of being caught
responded inaccurately.
• Fish consumers who purchased fish from illegal fish markets and believed
them to be commercial fish responded with lower consumption values.
• Surveys were not conducted in the native languages, and the details of the
survey were lost in translation when individuals had conversational English
skills only.
• Individuals surveyed relied heavily on fish for basic nutritional needs due to
economic necessity, or because of personal preference and/or cultural
traditions, and were afraid of restrictions that might jeopardize their family.
• Fishers understood the implications of the survey and responded inaccurately
out of pride.
• Surveys addressed only certain species of fish that were caught, yet fishers
caught and consumed numerous species of bottomfish.
• Questions were asked that made assumptions about the parts of fish
consumed when the whole fish, including organs, may have been consumed.
Each of these issues has been addressed in more than one recent fish consump-
tion survey in the past 2 years. Many fisheries resources and health officials
therefore believe that approaches that utilize community-level organizations
facilitate the survey process. This approach builds on the established trust
between the community organization and its members and enables surveyors to
develop a more accurate representation of fish consumption patterns.
Fish catch rates have also been used to estimate consumption rates, but varia-
tions in preparation methods, illegal resale of fish, and catching and preserving
fish for later consumption in other seasons and for extended families and friends
all add significantly to the uncertainty of these estimates. The duration of the
survey may include only times of high exposure or can be comprehensive and
address consumption rates year round to include variations in catch rates and
preservation and preparation methods.
B-10
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APPENDIX B
Some specific concerns have arisen over the use of license survey methods.
Performance exaggeration has been noted for sport fisher respondents, particu-
larly for individuals who associate fishing with prestige or who travel greater
distances to reach a particular fishing location. Nonresponse bias has also been
noted with surveys conducted on licensed fishers: typically, fishers who traveled
shorter distances to reach a fishing destination, or who fished less frequently or
consumed smaller quantities of fish, were less likely to respond to surveys than
were more frequent fishers. Consequently, consumption rates may have been
overestimated somewhat from surveys conducted in this manner.
B.2.3 Intake Patterns and Bolus Dose
When characterizing the consumption patterns of fishers, it is important to
consider the intake patterns. Patterns of exposure are critical to evaluating
potential health risks. As discussed in Section 2.4.3.2, toxicity is related to both
the overall exposure to a contaminant and the time over which the contaminant
is consumed. Exposure durations and exposure frequency are important factors
in estimating whether toxicity may occur. Consuming a few large meals over a
very short period (a bolus dose) may cause acute exposure health effects,
whereas consumption of the same total quantity spread over a month or year may
cause chronic exposure effects, or no effects at all.
Bolus dose exposure may pose significant risks to:
• Children who
consume greater quantities in relation to their body weight than adults
have greater susceptibility to some contaminants
have less capability to detoxify some contaminants.
• Pregnant women, if the contaminant is known to cause fetal damage following
prenatal exposure. Evidence from animal or human data presented in Section
5 shows that prenatal exposure to many of the target analytes may cause
damage to offspring.
• Persons with special susceptibilities due to illness (e.g., persons with kidney,
liver, or other diseases may be especially vulnerable to toxicants that attack
those systems).
The reader is urged to review the toxicity data provided in Section 5 for con-
taminants of interest in their areas to determine if there are population subgroups
requiring particular attention.
Fish consumption is often intermittent based on fish availability, cultural practices,
weather, and other factors. Determining whether a large intake is likely to occur
over a brief period of time is required to assess whether acute toxicity or develop-
mental toxicity may occur. It is important to obtain descriptive or quantitative infor-
mation on the timing of consumption over a calendar year.
B-11
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APPENDIX B
B.2.4 Calculation of Intake
When information is collected on both consumption patterns and contaminant
level, the contaminant exposure can be estimated. The contaminant exposure is
calculated using the fish consumption estimates for a specified time period (e.g.,
1 week, 1 month). The concentration of the contaminant in the fish (in milligrams
of contaminant per gram of fish) is multiplied by the amount of fish consumed (in
grams) during the time period to obtain the total contaminant exposure during that
time period (in milligrams). For example, if the contaminant concentration is 0.01
mg/g offish tissue, and 1,000 g offish are consumed in 1 month, then 0.01 mg/g
is multiplied by 1,000 g/mo to obtain a total exposure of 10 mg/mo.
To facilitate the risk assessment process, exposure is expressed in terms of the
daily average. The average daily exposure is calculated by dividing the total
amount of chemical contaminant ingested (in milligrams) during the specified
period by the number of days in the time period. For example, when data are
collected for a 1-month period, the following equation can be used to calculate
daily exposure:
average daily _ contaminant ingested over 1 month (mg/mo)
exposure (mg/d) days per month (d/mo) ' (°"1)
Although this equation uses 1 month as an averaging period, other averaging
periods could be used by changing the time periods in both the numerator and
denominator of the equation (e.g., 1 week).
Toxicity and risk values are expressed as intake in milligrams of chemical
contaminant per kilogram of body weight per day (mg/kg-d). To adapt the
exposure data to these units, the average daily exposure (in milligrams) is divided
by the body weight of the consumer (in kilograms):
average daily _ average daily exposure (mg/d)
intake (mg/kg-d) " body weight of consumer (kg) ' (D"2)
The most accurate body weight information is obtained directly from the local
population. Table 3-5 in Section 3 of this volume provides body weights for men,
women, and children of various ages from a national survey for use when local
data are not available.
To determine the potential for acute or prenatal toxicity, the total intake over a
short period of time (e.g., 3 days, 1 week) can be calculated. Depending on the
toxicity data being used, the time period of interest will vary (see Section 5 for
B-12
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APPENDIX B
chemical-specific information). The total intake is expressed as milligrams per
kilogram of body weight, as in the following equation:
total intake (mg/kg) = average daily intake (mg/kg-d)
x number of days (d) . ( ~ )
Information regarding the duration and periodicity of exposure is needed for both
determining potential risks and identifying the most appropriate consumption
limits. It should be described when exposure information is presented for use in
risk assessment.
B.3 FISH CONSUMPTION DATA FOR VARIOUS POPULATIONS
This section describes the results of fish consumption surveys. If state agencies
cannot conduct local surveys of fish consumption, these surveys can be used to
estimate fish consumption rates for the populations that an agency wishes to
target when issuing fish advisories. To use these data appropriately, it is important
to match the population surveyed in the reported studies as closely as possible
to the local fisher population. This section contains tables summarizing
consumption data for sport and subsistence fishers from studies conducted in
various regions of the United States. If a study is to be used as the basis for risk
assessment and setting advisory limits, agencies are strongly encouraged to
review the actual study data to determine its applicability to their local conditions.
Two categories of fisher survey data are discussed: sport fishers and subsistence
fishers. In these groups there is wide variability in consumption patterns. Although
the surveys are divided into these two categories for ease of presentation, these
two categories cannot be strictly defined. The results of many of these surveys
are summarized in Tables B-3 through B-6. They are presented by Region,
proceeding from east to west across the United States.
Tables B-3 and B-5 present consumption rate data for sport and subsistence
fishers, respectively. The tables list consumption in grams per day; however, it
should be noted that these values are estimates that are generally obtained by
recall, not strict log-keeping. In addition, surveys generally ask about the number
of meals eaten in a given time frame, but the size of these meals is generally
imprecisely estimated. In addition to quantitative data, information regarding the
types of fish included in the consumption rates is included with the consumption
rate, because it directly impacts the quantitative data presented in the rate tables.
These distinctions include
• Inclusion of freshwater fish, saltwater fish, or both
• Inclusion of sport and/or commercially caught fish.
B-13
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APPENDIX B
Survey methods used to collect the data reported in Tables B-3 and B-5 are listed
in Tables B-4 and B-6. The methods of conducting fish consumption surveys and
the reporting of information from these surveys may differ among studies and
many of the differences are highlighted in the survey methods tables.
Methods of averaging fish consumption information also differ among studies.
Some studies average the consumption rates over all individuals, regardless of
whether they ate fish, while other surveys average the information only for those
individuals who reported eating fish. For example, Cox et al. (1993) report
consumption rates averaged for the fish-eating population, whereas the Alabama
Department of Environmental Management (ALDEM, 1993) reports a rate
averaged for both the fish-consuming and nonconsuming populations. Although
some of the survey characteristics are noted in the tables, agencies should
consult the individual surveys to obtain the most complete descriptions of the
study and resulting consumption rates.
In addition to the studies of sport and subsistence fishers, national survey results
are discussed at the end of this section. In the absence of local data, national fish
consumption data may be used.
B.3.1 Sport Fishers
As noted previously, sport fishers differ with respect to their catch and consump-
tion habits. Some may fish for 1 week during a year or for several weekends each
year. Others may fish for much longer periods during a year or may fish year-
round. Surveys of the general sport fishing population may include those who
primarily fish for recreational purposes or eat fish for a small portion of the year
but may also include some individuals who eat fish as a main staple in their diets.
Fish consumption data obtained from sport fisher surveys are summarized in
Table B-3 and the survey methods used to collect the data are summarized in
Table B-4.
B-14
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APPENDIX B
Table B-3. Sport Fishers3 Consumption Data
Fisher Group
Alabama fishers1
Louisiana (coastal) fishers2
New York fishers3
New York (Hudson River)
fishers4
Michigan fishers5
Michigan fishers6
Michigan fishers7
Wisconsin fishers (10
counties)8
Wisconsin fishers (10
counties)8
Ontario fishers9
Los Angeles Harbor fishers10
Washington State
(Commencement Bay)
fishers11
Washington State (Columbia
River) fishers12
Maine fishers (inland
waters)13
Consumption Rates (g/d)
80th 90th 95th
Mean Median Percentile Percentile Percentile
45.8 50.7
65
28.1
40.9
14.5 30 62 80
18.3 =50
44.7
12.3 37.3
26.1 63.4
22.5
37 225
23 54
7.7
6.4 2.0 13 26
Fish Type
F+S, F+C
F+S, F+C
F+S, R+C
F+S, R
F+S, R
F+S, R+C
F, R
F, R
F, R+C
F, R
S, R
S, R
F+S, R+C
F, R
F = freshwater, S = saltwater, R = recreationally caught, C = commercially caught.
a Sport fishers may include individuals who eat sport-caught fish as a large portion of their diets.
SOURCES:
1 ALDEM (1993).
2 Dellenbargeretal. (1993).
3 Connelly etal. (1990).
4 Barclay (1993).
5 West etal. (1993).
6 West etal. (1989).
7 Uliimnhraw /1Q7R\
8 Fiore etal. (1989).
9 Cox etal. (1993).
10 Puffer etal. (1982).
11 Pierce etal. (1981).
12 Honstead etal. (1971).
13 Ebert etal. (1993).
B-15
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Table B-4. Sport Fishers3
Number
Fisher Group Surveyed
Alabama fishers1
Louisiana (coastal) fishers2
New York fishers3
New York (Hudson River)
fishers4
Michigan fishers5
Michigan fishers6
Michigan fishers7
Wisconsin fishers (10
counties)8
Ontario fishers9
Los Angeles Harbor fishers10
Washington State
(Commencement Bay)
fishers11
Washington State (Columbia
River) fishers12
Maine fishers (inland waters)13
NA= Not available.
a Sport fishers may include some
1,586
1,100
4,530
336
2,684
1,104
182
801
494
1,059
508
10,900
1,612
Contact Method/
Instrument
Onsite/personal
interview
Random/telephone
Fish license/mail/
followup by telephone
Onsite/personal
interview
Fish license/mail
Fish license/mail
Fish license/NA
Fish license/mail
Fish license/mail
Onsite/personal
interview
Fish license/personal
interview/followup by
telephone
Fish license/personal
interview
Fish license/mail/
followup by mail
individuals who eat fish as a large
Reporting
Method"
Log
Recall
Recall
Recall
Recall
Recall
Log
Recall
Recall
Recall
Recall
Recall
Recall
portion of their
b Respondents recorded consumption information in a log or recalled consumption
c Catch: Original data from catch
SOURCES:
1 ALDEM(1993).
2 Dellenbargeret al. (1993).
rates extrapolated to consumption
3
Connelly et al.
(1990).
Survey Description
Catch vs.
Consumption0
Catch
Consumption
Catch
Consumption
Consumption
Consumption
Catch
Consumption
Consumption
Catch
Catch
Consumption
Consumption
diets.
Individual vs.
Household
Individual
Household
Individual
NA
Household
Household
Individual
Individual
Individual
Individual
Individual
Household
Individual and
household
Data Available
Age, ethnicity, income,
region, sex
Age, education, ethnicity,
income, other
Age, income, region
NA
Age, education, ethnicity,
income, region, sex
Age, education, ethnicity,
income, region, sex
NA
Age, education, ethnicity,
region, sex
Age, region, sex
Age, ethnicity
NA
NA
NA
Duration
12 mo
1 mo
12 mo
NA
12 mo
6 mo
24 mo
NA
Summer, fall
12 mo
Summer, fall
12 mo
12 mo
information during interview.
rates. Consumption: Data obtained on consumption patterns.
4 Barclay (1993). 6 Westetal.
5 West et al.
(1993).
(1989).
>
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APPENDIX B
Table B-5. Subsistence Fishers3 Consumption Data
Consumption Rates (g/d)
Fisher Group
Great Lakes tribes1
Columbia River tribes2
High-end Caucasian consumers on
Lake Michigan3
Native Alaskan adults4
Mean
351
58.7
48b
27c
109
95th percentile Max
1,426
170
144
132
Fish Type
F
F
F
F
F+S
F = fish, S = shellfish.
a Subsistence fishers include individuals who may eat sport-caught fish at high rates but do not subsist on fish as a
large part of their diet.
b Data from 1982 survey of fish eaters.
c Data from 1989 survey of fish eaters.
SOURCES:
1 Kmiecikand Ngu (1994).
2 CRITFC(1994).
3 Hovingaetal. (1992, 1993).
4 Nobmanetal. (1992).
B.3.2 Subsistence Fishers
Subsistence fishers consume fish as a major staple of their diet. These fishers
rely on fish to meet nutritional needs, as an inexpensive food source, and, in
some cases, because of their cultural traditions. Subsistence fishers often have
higher consumption rates than other fisher groups; however, consumption rates
vary considerably among subsistence fishers. Consequently, generalizations
should not be made about this fisher group. If studies contained in this section are
used to estimate exposure patterns for a subsistence population of concern, care
should be taken to match the dietary and population characteristics of the two
populations as closely as possible.
Subsistence fishers include a wide variety of people who differ in many respects.
This section is not suggesting that similarities exist between populations, other
than in their consumption of a relatively large quantity of fish. Information is
provided below on some qualitative characteristics of specific subsistence
population groups.
B-17
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DO
Table B-6. Subsistence Fishers3 Survey Description
Fisher Typeb
Great Lakes tribes1
Columbia River
tribes2
High-end Caucasian
consumers on Lake
Michigan3
Native Alaskan
adults4
NA = Not available.
Number Contact Method/ Reporting Catch vs. Individual
Surveyed Instrument Method" Consumption0 vs. Household
69 Tribe/mail Recall Consumption Individual
717 Tribe/random/personal Recall Consumption Individual
interview
115 Otherd/personal Recall Consumption Individual
interview
351 Tribe/random/personal Recall Consumption Individual
interview
a Subsistence fishers include individuals who may eat sport-caught fish at high rates but do not subsist on fish as a large
Data Duration
Available (months)
NA 2
Age, ethnicity, 12
region, sex
Age, sex, 7
education, other
Age, ethnicity, 18
sex, other
part of their diets.
b Respondents recorded consumption information in a log or recalled consumption information during interview.
c Catch: Original data from catch rates extrapolated to consumption rates. Consumption: Data obtained on consumption
d Fishers identified in a
SOURCES:
Michigan Department of Health study in 1982.
patterns.
1 Kmiecik and Ngu (1994).
2 CRITFC(1994).
3 Hovinga et al. (1992,
4 Nobman et al. (1992)
1993).
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APPENDIX B
Subsistence fishers may consume different types or portions of fish than sport
fishers (e.g., organs, whole fish), although individual tastes will vary. Their
consumption patterns in this regard may result in greater exposure to con-
taminants. For example, many Asian-American subsistence fishers eat raw fish,
liver, hepatopancreas, kidneys, brains, and eyes of bottom-dwelling fish such as
carp and catfish that bioaccumulate more toxicants due to the scavenging habits).
They may use whole fish in soup stocks and consume seaweed and other aquatic
species that may contain the same contaminants as fish. Fish advisory programs
have only recently begun to address concerns associated with this subpopulation,
and some studies are underway to evaluate consumption patterns. Current
information is primarily qualitative; however, differing patterns have been identified
among the populations considered: Laotians, Hmong, Cambodian, and
Vietnamese (Allbright, 1994; Cung, 1994; Den, 1994; Lorenzano, 1994; Nehls-
Lowe, 1994; Pestana, 1994; Shubat et al., 1996; University of Wisconsin Sea
Grant, 1994; Young, 1994).
Native American groups in some areas include fish extensively in their cultural,
ceremonial, and dietary patterns. Many of the surveys of Native American groups
indicate a high fish consumption rate. Most of the study information is recent and
many studies are still ongoing.
Rural fishers make up a large segment of subsistence fishers. For example, more
than half the noncommercial fishing in Idaho is conducted in Washington County,
Idaho. Within Washington County, a community considered by some researchers
to be subsistence fishers is located in the area surrounding Brownlee Reservoir,
a major fishing location. The local community has a high unemployment rate, with
over 40 percent of the population on public assistance. The sport and subsistence
fishers in the area often catch 100 to 300 Ib of crappies during a fishing trip and
freeze much of the catch for year-long consumption. Many fishers are dependent
on fish as a major source of protein for themselves and their families. Fishing
activities also bring needed economic resources to the area. However, elevated
pollutant levels have been found in the reservoir. Community leaders have
concerns regarding tradeoffs between fish advisories developed to reduce health
risks and the negative economic and nutritional impacts the advisories might have
on the fisher population (Richter and Rondinelli, 1989).
Several surveys evaluating the consumption patterns of subsistence fishers have
been initiated in the past several years. Some of these have been completed and
many more are currently being carried out, with results expected in the near
future. Although many of these surveys provide only a range of consumption
rates, a great deal of qualitative information has been gained through these
surveys, both about the individual populations that were studied and about
effective survey methods for different groups of subsistence fishers. The
consumption rates reported by these surveys are presented in Table B-5 and the
survey methods used to collect the data are summarized in Table B-6.
B-19
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APPENDIX B
B.3.3 General Population
For the purposes of risk assessment or risk management, the consumption rates
derived from national surveys can provide a useful picture of the distribution of
fish consumption for the U.S. population. However, since sport and subsistence
fishers generally have higher consumption rates than the national rates, the
distributions for these groups will differ. That is, the point estimates of the mean
and upper percentiles of fish consumption will generally be higher for the sport
and subsistence fishers than for the general U.S. population. National survey data
are the least preferred for use in developing local advisories.
Fish consumption data from three national studies are reported in Table B-7. The
details of the survey methods used in these studies are summarized in Table B-8.
Note that two of the three studies (National Purchase Diary [NPD] and Market
Facts) were conducted more than 20 years ago. Also, study results conflict in
some respects. For example, the NPD study found the lowest consumption rate
in New England, and the Market Facts study found the highest rates in New
England. There is also concern that the reported rates in these dated studies do
not reflect current consumption patterns.
B.3.4. Sensitive Subpopulations
States with consumption rate information specific to sensitive subpopulations
(e.g., women of reproductive age and children) may wish to use such information
when assessing exposure. For example, a recent study was conducted to
determine fish consumption patterns among the Umatilla, Nez Perce, Yakama,
and Warm Springs Tribes of the Columbia River Basin in Washington and Oregon
(CRITFC, 1994). This study found that adults in these four tribes consume an
average of 58.7 g/d and that children (5 years and younger) from these four tribes
consumed 19.6 g/d. Mean fish consumption was more than nine times higher
among adults and over three times higher among children in these tribes than for
adults in the general population (assuming a consumption rate of 6.5 g/d). Many
of the contaminants examined in Section 5 of this volume have develop-mental
effects of particular concern to women of reproductive age and children.
If data are available for only the general population, however, the consumption
rates for the populations of interest may be calculated by using values for meal
size and body weights specific to those subgroups using the methods described
in Section 3 of this volume. In cases where studies do not separate consumption
rates by age and gender, an exposure assessment based on these rates would
reflect exposure to the general population only.
Population size estimates may need to be adjusted to include family members of
fishers who share their catch. While children may not constitute a large fraction
of fishers, they may be exposed by eating fish that their parents or older siblings
catch. Site-specific data on family size can be used to make this estimate, if
B-20
-------
APPENDIX B
Table B-7. National Studies Consumption Data
Consumption Rates (g/d)
Population
US1
US2
us2
us3
us4
us4
Mean
6.6
6.5
14.3
16.7
20.1
5.9
90th Percentile
NA
NA
NA
NA
70.1
15.9
95th Percentile
47.3
NA
41.7
NA
102.0
40.0
99th Percentile
NA
NA
NA
NA
173.2
107.6
Fish Type
F+E, C+R
F+E, C+R
F+S, C+R
F+S, C+R
F+S+E, C+R
F+E, C+R
F = Freshwater, S = Saltwater, E = Estuarine, C = Commercial, R = Recreational.
SOURCES:
1 Continuing Survey of Food Intake by Individuals (CSFII) conducted by USDA(1991).
2 National Purchase Diary (NPD) Fish Consumption Survey (as cited in Javitz, 1980; Rupp et al., 1980).
3 Market Facts Survey (as cited in Javitz, 1980).
4 Continuning Survey of Food Intake by Individuals (CSFII) conducted by USDA, 1988, 1990, 1991, U.S. EPA
(1998b).
available. In the absence of these data, U.S. census data on average family size
can be used.
Other susceptible subpopulations among the fisher populations should be con-
sidered as well. The presence of these groups will depend on local demographics
and the nature of the contaminants present in fish. Section 5 of this volume
provides information on especially susceptible subgroups for many of the target
analytes. Some chemical contaminants interfere or act synergistically with
Pharmaceuticals; others attack particular organ systems and may cause people
with related illnesses to be at elevated risk. Information on any susceptible
subgroup should be considered both in estimating risks and establishing health-
based exposure limits.
B-21
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Number
Population Surveyed
US1 11,912
US2 23,213
US3 4,864
US4 11,912
Table B-8
Contact
Method/
Instrument
Census/personal
interview
Census/NA
Census/NA
Census/personal
interview
. National Studies Survey Description
Reporting
Method3
Log/recall
Log
Log
Log/recall
Catch vs.
Consumption13
Consumption
Consumption
Consumption
Consumption
Individual
vs. Data
Household Available Duration
Individual Age, sex 12 mo
(3 d recall/
person)
Household Age, sex, 12 mo
region
Household Education, 12 mo
ethnicity,
income
Individual Age, sex 12 mo
(3 d recall/
person)
NA = Not available.
a
b
Respondents recorded consumption information in a log
Catch: Original data from catch
or recalled consumption information during interview.
rates extrapolated to consumption rates.
Consumption: Data obtained on consumption patterns.
SOURCES:
1
2
3
4
Continuing Survey of Food Intake by Individuals (CSFII)
National Purchase Diary (NPD)
conducted by USDA (1991).
Survey (as cited in Javitz, 1980; Rupp et
al., 1980).
Market Facts Survey (as cited in Javitz, 1980).
Continuing Survey of Food Intake by Individuals (CSFII) conducted by USDA, 1989,1990, 1991,
U.S. EPA(1998b).
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APPENDIX B
B.4. CONSUMPTION SURVEY DATA ORGANIZATION
In assembling the exposure data, it is most appropriate to build a population
exposure database in the form of data groupings for each waterbody and
population subgroup (e.g., population consumption characteristics for individuals
living around or using a particular lake, river, etc.). Because most contamination
data are maintained for specific waterbodies, they serve as a natural unit for
evaluating exposure.
Further subdividing of a population may be necessary, depending on population
size and the area being considered. If a large or diverse population of concern
(e.g., a city or large geographic area) is to be evaluated, subgroups within the
population of interest may need to be identified. These subgroups, which may
have higher than average exposures, can include groups of subsistence fishers
or sport fishers known to fish in contaminated waters. If attention is focused on
smaller groups (e.g., sport fishers at a single lake, subsistence fishers from a
particular tribe), further subdividing the population into subgroups may not be
necessary for purposes of evaluating exposures.
A template is provided in Section 2, Table 2-4, of this volume on which exposure
data may be entered. It is located in that section because risk managers are
encouraged to evaluate other aspects of exposure in addition to consumption
patterns. These factors include exposure modifications that may be associated
with fish cleaning (skinning and trimming) and cooking fish procedures (discussed
in Appendix C) and additional exposures to the contaminant of concern that may
arise from other sources such as air, water, other foods, and soil (discussed in
Section 2.4.5.6 of this volume).
B.5 REFERENCES
ALDEM (Alabama Department of Environmental Management). 1993. Estimation
of Daily Per Capita Freshwater Fish Consumption of Alabama Anglers.
Prepared by Fishery Information Management Systems, Inc., and the
Department of Fisheries and Allied Aquacultures, Auburn University, AL.
Allbright, Kelly. 1994. Minnesota Department of Health, Division of Environmental
Health. Personal communication with Abt Associates, May 27, May 31, July
28.
Anderson, H.A., and J.F. Amrhein. 1993. Protocol for a Uniform Great Lakes
Sport Fish Consumption Advisory. Prepared for the Great Lakes Advisory
Task Force. May.
Barclay, Bridget. 1993. Hudson River Angler Survey. Poughkeepsie, NY: Hudson
River Sloop Clearwater, Inc.
B-23
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APPENDIX B
Carlson, G. 1994. Comments on Volume 2, Risk Assessment and Fish Consump-
tion Limits (first edition) from the Missouri Department of Health. April 22.
Connelly, NA., T.L. Brown, and B.A. Knuth. 1990. New York Statewide Angler
Survey 1988. New York State/Department of Environmental Conservation,
Division of Fish and Wildlife, Albany, NY. 158 pp.
Cox, C., A. Vaillancourt, and A. Hayton. 1993. The Results of the 1992 Guide to
Eating Ontario Sport Fish. PIBS Questionnaire 2593E. Ministry of
Environment and Energy, Ontario, Canada. November.
CRITFC (Columbia River Inter-Tribal Fish Commission). 1994. A Fish
Consumption Survey of the Umatilla, Nez Perce, Takama, and Warm Springs
Tribes of the Columbia River Basin. CRITFC Technical Report #94-3.
Portland, OR.
Cung, Josee. 1994. Minnesota Department of Natural Resources. Southeast
Asian Outreach Project. Personal communication with Abt Associates. July
28.
Dellenbarger, L, A. Schupp, and B. Kanjilal. 1993. Seafood Consumption in
Coastal Louisiana. Louisiana Department of Environmental Quality.
Den, Arnold. 1994. Senior Science Advisor, U.S. Environmental Protection
Agency Region 9. Personal communication with Abt Associates. July 21, July
28.
Dourson, M.L., and J.M. Clark. 1990. Fish consumption advisories: Toward a
unified, scientifically-credible approach. Reg. Tox. Pharmacol. 12(2):161-178.
Ebert, E.S., N.W. Harrington, K.J. Boyle, J.W. Knight, R.E. Keenan. 1993.
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Fiore, B.J., H.A. Anderson, L.P. Hanrahan, L.J. Olson, and W.C. Sonzogni. 1989.
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Honstead, J.F., T.M. Beetle, and J.K. Soldat. 1971. A Statistical Study of the
Habits of Local Fishermen and Its Application to Evaluation of Environmental
Dose. Battelle Pacific Northwest Laboratories, Richland, WA.
Hovinga, M.E., M.F. Sowers, and H.E.B. Humphrey. 1992. Historical changes in
serum PCB and DDT levels in an environmentally exposed cohort. Arch. Env.
Contam. Toxicol. 22 (4): 362-366.
B-24
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APPENDIX B
Hovinga, M.E., M. Sowers, and H.E.B. Humphrey. 1993. Environmental exposure
and life-style predictors of lead, cadmium, PCB, and DDT levels in Great-
Lakes Fish Eaters. Arch. Env. Hea/tf?48(2):98-104. May
Humphrey, H. 1976. Evaluation of Changes of the Level of Polychlorinated
Biphenyls (PCBs) in Human Tissues. Final report on FDA contract 223-73-
2209. Michigan Department of Public Health, Lansing.
Javitz, Harold. 1980. Seafood Consumption Data Analysis, Final Report. SRI
International. Prepared for the U.S. Environmental Protection Agency, Office
of Water Regulations and Standards, Task 11, EPA Contract 68-01-3887.
Kmiecik, Neil, and H.H. Ngu. 1994. Survey of Tribal Spearer: Mercury Concerns.
Great Lakes Fishing Memorandum. April 20.
Lorenzano, R. 1994. U.S. Environmental Protection Agency Region 10. Personal
Communication with Abt Associates. July 28.
Minnesota Department of Health. 1992. Minnesota Fish Consumption Advisory.
Minneapolis, MN. May.
Missouri Department of Health. 1992. 1992 Fish Consumption Advisory. Jefferson
City, MO. May.
Nehls-Lowe, Henry. 1994. Wisconsin Department of Natural Resources. Personal
communication with Abt Associates. July 29.
Nobman, E.D., T. Byers, A.P. Lanier, J.H. Hankin, M.Y. Jackson. 1992. The diet
of Alaska native adults. Am J Clin Nutr 55(5): 1024-32.
Pestana, Edith. 1994. Connecticut Commissioner's Office of the Department of
Environmental Protection, Section of Environmental Justice. Personal
communication with Abt Associates, May 18.
Pierce, R.S., D.T. Noviello, andS.H. Rogers. 1981. Commencement Bay Seafood
Consumption Report. Preliminary Report. Tacoma-Pierce County Health
Department, Tacoma, WA.
Puffer, H.W., S.P. Azen, M.J. Duda, and D.R. Young. 1982. Consumption Rates
of Potentially Hazardous Marine Fish Caught in the Metropolitan Los Angeles
Area. EPA 600/3-82-070. U.S. Environmental Protection Agency,
Environmental Research Laboratory, Corvallis, OR.
Richter, B.S., and R. Rondinelli. 1989. The Relationship of Human Levels of Lead
and Cadmium to the Consumption of Fish Caught In and Around Lake Coeur
d'Alene, Idaho. Final Report. Technical Assistance to the Idaho State Health
Department and the Indian Health Service, Boise, ID.
B-25
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APPENDIX B
Rupp, Elizabeth, F.L. Miller, and I.C.F. Baes III. 1980. Some results of recent
surveys offish and shellfish consumption by age and region of U.S. residents.
Health Physics 39:165-175.
Shubat, P. 1993. Minnesota Department of Health. Conversation with Abt
Associates. August 25.
Shubat, P.J., K.A. Raatz, and R.A. Olson. 1996. Fish consumption advisories and
outreach programs for Southeast-Asian immigrants. Toxicol. Ind. Health 12
(3-4):427-434.
University of Wisconsin SeaGrant. 1994. Personal communication with Abt
Associates. May 27.
USDA(U.S. Department of Agriculture). 1988, 1990, 1991. Continuing Survey of
Food Intakes by Individuals Data and Documentation. Human Nutrition
Information Service, Hyattsville, MD.
U.S. EPA (Environmental Protection Agency). 1988. Region V Risk Assessment
forDioxin Contaminants. Chicago, IL.
U.S. EPA (Environmental Protection Agency). 1992. Consumption Surveys for
Fish and Shellfish: A Review and Analysis of Survey Methods. Office of
Water, Washington, DC.
U.S. EPA (Environmental Protection Agency). 1995. Guidance for Assessing
Chemical Contaminant Data for Use in Fish Advisories. Volume 1: Fish
Sampling and Analysis, Second Edition. Office of Science and Technology,
Office of Water, Washington, DC.
U.S. EPA (Environmental Protection Agency). 1998a. Guidance for Conducting
Fish and Wildlife Consumption Surveys. EPA-823-B-98-007. Office of Water,
Washington, DC. November.
U.S. EPA (Environmental Protection Agency). 1998b. Daily Average Per Capita
Fish Consumption Estimates Based on the Combined USDA 1989, 1990, and
1991 Continuing Survey of Food Intakes by Individuals (CSFII). Volume I:
Uncooked Fish Consumption National Estimates. Office of Science and
Technology, Washington, DC. Submitted by Sciences Applications
International Corporation, Environmental Health Sciences Group. March.
U.S. FWS (Fish and Wildlife Service). 1993. National Survey of Fishing, Hunting,
and Wildlife Associated Recreation. Washington, DC.
B-26
-------
APPENDIX B
West, P.C. M.J. Fly, R. Marans, and F. Larkin. 1989. Michigan Sports Anglers
Fish Consumption Survey, Supplement I, Non-Response Bias and
Consumption Suppression Effect Adjustments. Technical Report No. 2.
Natural Resource Sociology Research Lab, School of Natural Resources,
University of Michigan, Ann Arbor.
West, P.C., M.J. Fly, R. Marans, and F. Larkin. 1993. 1991-92 Michigan Sport
Anglers Fish Consumption Study. Final Report to the Michigan Great Lakes
Protection Fund, Michigan Department of Natural Resources, Lansing, Ml.
Technical Report No. 6. School of Natural Resource Sociology Research Lab,
University of Michigan. May.
Young, Pat. 1994. U.S. Environmental Protection Agency Region 9. Personal
communication with Abt Associates, July 28.
B-27
-------
APPENDIX C
DOSE MODIFICATIONS DUE TO FOOD
PREPARATION AND COOKING
-------
-------
APPENDIX C
APPENDIX C
DOSE MODIFICATIONS DUE TO FOOD PREPARATION AND COOKING
C.1 DOSE MODIFICATIONS OF FISH CONTAMINANT EXPOSURE
Fish preparation and cooking procedures can modify the amount of contaminant
ingested by fish consumers. Consequently, exposure and dose are modified.
Incorporating a dose modification factor into the exposure equation to account for
loss of chemical contaminants from fish tissue during preparation and cooking
requires two types of information:
• Methods used by fish consumers to prepare (trimming, skinning) and cook
(broiling, baking,, charbroiling, canning, deep frying, pan frying, microwaving,
poaching, roasting, salt boiling, smoking) their catch.
• The extent to which a particular contaminant concentration is likely to be
decreased by these culinary methods.
To adjust contaminant concentrations appropriately, the dose modification factors
must be matched to the type of sample from which the fish contaminant
concentration was measured. For example, it would be inappropriate to apply a
dose modification factor for removing skin if the contaminant concentrations in
the fish were based on the analysis of a skin-off fillet. To select the correct
approach for evaluating exposure, information on both the distribution of
chemicals in fish tissue and alterations due to food preparation and cooking must
be used. The modified contaminant concentration (based on preparation and
cooking losses) is used to modify the exposure estimates used in the risk
equations. This information is also useful in development of fish advisories and
risk communication activities.
C.1.1 Contaminant Distribution in Fish Tissues
Chemical contaminants are not distributed uniformly in fish. Fatty tissues, for
example, will concentrate organic chemicals more readily than muscle tissue.
Muscle tissue and viscera will preferentially concentrate other contaminants. This
information has important implications for fish analysis and for fish consumers.
Depending on how fish are prepared and what parts are eaten, consumers may
have significantly differing exposures to chemical contaminants. This section is
meant as an overview; states should consult primary research studies for more
information. In general, contaminant concentrations differ among
C-3
-------
APPENDIX C
Fatty tissues, muscle tissue, and internal organs
Different species of fish
Different age or size classes of fish
Type of chemical contaminant present in the fish.
C.1.2 Fish Tissue Types
Lipophilic chemicals accumulate mainly in fatty tissues, including the belly flap,
lateral line, subcutaneous and dorsal fat, and the dark muscle, gills, eyes, brain,
and internal organs. Some heavy metals, such as cadmium, concentrate more in
the liver and kidneys. Muscle tissue often contains lower organic contaminant
concentrations than fatty tissues (Great Lakes Sport Fish Advisory Task Force,
1993), but contains more mercury, which binds to muscle proteins (Minnesota
Department of Health, 1992).
Many people remove the internal organs before cooking fish and trim off fat and
skin before eating, thus decreasing exposure to lipophilic and other contaminants.
Removing the fat, however, will not decrease exposure to other contaminants,
such as mercury, that are concentrated in muscle and other protein-rich tissues
(Gutenmann and Lisk, 1991; Minnesota Department of Health, 1992). Concentra-
tions of mercury have been shown to be higher per gram of fillet in skin-off than
in skin-on fillets contaminated with mercury (Gutenmann and Lisk, 1991). Certain
populations, including some Asian-Americans and Native American groups, eat
parts of the fish other than the fillet and may consume the whole fish. Recipes
from many cultures employ whole fish for making soups or stews. As a result,
more of the fish contaminants are consumed.
States should take preparation methods of local fisher populations
into account when assessing exposure levels and when assessing
whether use of a dose modification factor is appropriate for their
target fish-consuming population.
C.1.3 Fish Species
Fish accumulate contaminants from the water column, from suspended sediment
and organic matter in the water, and from their food. Depending on their
propensity to bioaccumulate contaminants (largely a function of their feeding
habits, ability to metabolize contaminants, and fat content), different fish species
living in the same area may contain very different contaminant concentrations.
Due to biomagnification, higher trophic level species are more likely to have
higher contaminant concentrations. The tissues of the top predators can contain
contaminant levels exceeding those in ambient water or sediments by several
orders of magnitude.
Where a fish feeds in the water column also determines its relative bio-
accumulation potential. Bottom feeders, such as carp or catfish, are exposed to
C-4
-------
APPENDIX C
more sediments than are fish that feed in mid-water or near the surface of the
water column. Bottom feeders, therefore, have a tendency to accumulate more
of the dense, hydrophobic contaminants, such as chlordane or polychlorinated
biphenyls (PCBs), that are adsorbed to the sediment particles. In addition, fish
species vary widely in their fat content. Fish low in fat, such as bass, sunfish,
crappies, yellow perch, and walleyes, are less likely to accumulate lipophilic
contaminants than fattier fish such as bluefish, rainbow trout, lake trout, some
salmon, catfish, and carp. Even within the same species, great differences in fat
content may occur. Zabik et al. (1996) reported the average fat content of Lake
Michigan lean lake trout (Salvelinus namaycush namaycush) was 9.1 percent,
which was significantly lower than that of the fat Lake Superior siscowets
(Salvelinus namaycush siscowet) (20.5 percent). Aquatic organisms also differ
in their abilities to metabolize and excrete contaminants. For example, one study
found fish more readily able to metabolize benzo[a]pyrene than shrimp, amphipod
crustaceans, and clams, respectively (U.S. EPA, 1995a). The ability to break
down and excrete chemical contaminants may also differ among fish species.
This differential accumulation of contaminants produces very different exposure
levels for individuals eating different species of fish. An individual who eats
primarily fatty fish species will receive higher exposures of organic chemical
contaminants than an individual who eats primarily leaner fish species. Thus,
states should consider multiple species exposure in their decision to issue fish
consumption advisories.
C.1.4 Fish Size or Age Class
Larger size classes of fish within the same species generally contain higher
concentrations of bioaccumulative contaminants, especially the more persistent
chemicals such as mercury, DDT, PCBs, and toxaphene (Gutenmann et al. 1992;
U.S. EPA, 1995a). Because larger fish are older, they have had more time to
accumulate chemicals from their food and they are more likely to catch larger
prey, which themselves have had a longer time to bioaccumulate chemicals
(Minnesota Department of Health, 1992). Older fish also concentrate more
contaminants in their muscle tissues, which are fattier than muscle tissue in
younger fish, particularly along the backbone and lateral lines (Kleeman et al.,
1986a). States may choose to issue size-specific consumption advisories and/or
explain this correlation of increasing contaminant residues in larger fish within a
given species in their public education efforts.
C.1.5 Chemical Contaminants
Many of the target chemicals examined in this guidance series are lipophilic and
accumulate in the fatty tissues. Some contaminants (and their congeners)
bioaccumulate in fish more readily than others or are more resistant to
metabolism and excretion once accumulated than others (Bruggeman et al., 1984;
Stern et al., 1992). Thus, fish exposed to the same concentrations of a
contaminant may accumulate different levels of contaminants in their tissues
C-5
-------
APPENDIX C
based on their ability to bioaccumulate the contaminant directly from solution or
via preconcentration on prey species coupled with their ability to metabolize and
excrete the contaminant.
States may wish to use this chemical-specific information on distribution of
contaminants in fish tissues to assess whether a local population may be
exposed unreasonably to a given contaminant, due to particular eating habits
such as eating only one species of fish, eating specific parts (whole fish or
organs) of the fish, or eating fish species with a high fat content in contrast to
eating leaner species.
C.1.5.1 Heavy Metals-
Several studies indicate that mercury, cadmium, and selenium bind to different
tissues in fish than do organochlorines. Mercury, for example, binds strongly to
proteins, thereby concentrating in muscle tissues of fish (Gutenmann and Lisk,
1991; Minnesota Department of Health, 1992). Mercury also concentrates in the
liver and kidneys, though at generally lower rates (Harrison and Klaverkamp,
1990; Marcovecchio et al., 1988). Thus, trimming and gutting can actually result
in a greater average concentration of mercury in the remaining fillet tissues
compared with the concentration in the whole untrimmed fish proteins, thereby
concentrating in muscle tissues offish (Gutenmann and Lisk, 1991).
Cadmium concentrates largely in the liver, followed by the kidneys and gills, and
less so in the muscle tissue (Harrison and Klaverkamp, 1990; Marcovecchio etal.,
1988; Norey et al., 1990), indicating that cadmium concentrations could be
decreased by trimming and gutting fish before consumption.
Selenium was shown to concentrate in both the liver and muscle tissues at similar
rates (Harrison and Klaverkamp, 1990). Consumers would be likely to receive a
lower exposure if they consumed a fillet only rather than consuming the whole fish
(including fillet tissue and the liver tissue).
C.1.5.2 Organochlorines—
Organochlorine pesticides, PCBs, dioxins/furans tend to concentrate in fatty
tissues (Armbruster et al. 1989; Branson et al., 1985; Bruggeman et al. 1984;
Gutenmann et al. 1992; Kleeman et al., 1986a, 1986b; Ryan et al., 1983; Skea
et al., 1979; Sanders and Hayes 1988; U.S. EPA, 1995a ). Many of these
compounds are neither readily metabolized nor excreted and thus tend to
biomagnify through the food web ( Gardner and White, 1990; Lake et al., 1995;
Metcalf and Metcalf, 1997; Muir et al., 1986; Niimi and Oliver, 1989; Oliver and
Niimi, 1988; U.S. EPA, 1995a). Because different fish species store fat differently,
contain different amounts of body fat, and metabolize these compounds at
slightly different rates, each species will also concentrate organochlorine-based
contaminants somewhat differently. In general, however, trimming away fatty
C-6
-------
APPENDIX C
tissues, including the skin, are the most effective ways to reduce exposure to
these chemicals.
C.1.5.3 Other Contaminants—
The other chemicals examined in this exposure assessment (organophosphate
pesticides and oxyfluorfen) have also been found to bioaccumulate in fish, but to
a much lower extent than the organochlorine pesticides. Little information is
available, however, on the distribution of these chemicals in specific fish tissues.
After feeding chlorpyrifos to channel catfish in a laboratory study, the highest
concentrations were found in the liver tissue, while less than 5 percent of the dose
was found in muscle tissue (Barren et al., 1991). No information was located on
the tissue distribution of any of the other organophosphates in feral fish
populations. Organophosphates as a group are lipophilic and would be expected
to distribute to body fat like the organochlorine compounds. However, the
organophosphates are much less persistent in both the environment (U.S. EPA,
1995a) and in aquatic organisms because these compounds are vulnerable to
hydrolysis in water and to metabolic breakdown by esterases.
C.2 ESTIMATING DOSE MODIFICATION BASED ON PREPARATION METHODS
This section presents data on the effects of various preparation methods on
contaminant concentrations in fish tissue. In the absence of specific data on fish
preparation methods, the U.S. Environmental Protection Agency (EPA)
recommends using fillets as the standard sample type for analyzing chemical
contaminants. Readers are referred to Volume 1, 3rd edition, of this series for a
more complete discussion of sample analysis (U.S. EPA, 1999). The sample type
should consist of the portion of the individual organism commonly consumed by
the general fish-consuming population or a specific target population of concern
(e.g., pregnant or nursing women, young children, recreational or subsistence
fishers). EPA recommends analyzing skin-on fillets (including the belly flap) for
most scaled finfish. Conversely, skin-off fillets may be more appropriate for target
species without scales (e.g., catfish). State or local agencies, however, are
advised to select the sample type most appropriate for each target species based
on consumption patterns of local populations and should sample the whole body
of the fish if a local target population typically consumes whole fish. Following
these guidelines, states may have concentration data from fillet samples with skin-
on, fillet samples with skin-off, or from whole fish.
When states have data on the preparation methods of the target fish-consuming
populations, appropriate dose modification factors from these studies can be used
to adjust assumed fish chemical contaminant concentrations. Without food
preparation data, however, states should not assume that specific preparation
methods are employed, since fish preparation and cooking techniques frequently
vary among individuals and often depend on the type of fish consumed. As noted
earlier, many groups known to consume large quantities offish, including Native
American and Asian American fishers, often consume most of the whole fish and
C-7
-------
APPENDIX C
may do very little trimming. Consequently, assuming a dose reduction in chemical
contaminants based on fillet samples may lead to an underestimate of the
exposure and risk for these groups that consume whole fish.
EPA recommends the use of dose modification factors for setting
health-based intake limits only when data on local methods of prepara-
tion and their impact on contaminant concentrations are available.
EPA recommends that all fish advisories emphasize the importance of skinning
and trimming fish (including gutting) and certain ways of cooking as effective
means to minimize the risks from chemical contaminant residues in fish tissue. To
achieve the best results, all three techniques should be used together. States are
encouraged to include illustrations in their fish advisories showing the location of
fatty tissue in fish and describing the parts of of the fish tissue to be trimmed.
This type of information could be provided to fish consumers as part of a fish
advisory program through risk communication efforts. Further information on risk
communication is included in Volume 4 in this series of guidance documents (U.S.
EPA, 1995b).
The degree of preparation-related reduction in contaminant concentration
depends on
• Fish species and size (age class)
• Chemical contaminant residues present
• Specific food preparation and cooking techniques used.
Consumer concern about the presence of toxic chemicals in fish has focused
research on quantitating the effects of processing and cooking on the possible
reduction of chemical contaminant levels in fish. Several generalizations about
specific food preparation and cooking techniques can be made based on several
detailed studies conducted using primarily Great Lakes fish.
• Trimming fish is an important consideration in reducing the levels of PCBs
and other organochlorine pesticides ingested by consumers (Hora 1981;
Sanders and Haynes, 1988; Zabiketal., 1995b; ZabikandZabik, 1996). For
example, in a recent study, raw skin-off fillets had an average of 50 percent
of the residues found in raw skin-on fillets. The skin-off fillets had both the
belly flap and the lateral line and its associated fat trimmed off, while the skin-
on fillet had only the belly flap removed. Zabik et al. (1995b) also established
that this contaminant reduction was carried over to cooked fillets.
• Cooking methods that allow the separation of the cooked muscle from the
skin (pan frying, poaching, broiling, baking) reduce the amount of chemical
contaminants the consumer would ingest over such cooking methods as
deep frying where both the skin and cooked muscle are consumed together
(Zabiketal., 1995a).
C-8
-------
APPENDIX C
• As a cooking process, smoking resulted in significantly greater reductions (40
to >50 percent) of organochlorine pesticides (DDT,DDE, ODD, chlordane
complex, HCB, dieldrin, heptachlor epoxide, toxaphene), total PCBs, and
dioxin residues (TCDD) than other cooking methods (baking, charbroiling,
salt boiling, deep fat frying, canning) tested, but polynuclear aromatic
compounds (PAHs) showed significant formation during the smoking process
especially in fish species with higher body fat levels (siscowet) (Zabik et al.,
1996).
• For dioxins, several organochlorine pesticides, and PCBs, increasing the
internal temperature of the cooked fish from 60 to 80 °C (Stachiw et al., 1988),
increasing the surface area exposed to the cooking process by scoring the
fillets (Stachiw et al.,1988: Zabik et al., 1994), or increasing the cooking time
or cooking temperature enhances the loss of contaminant residues in the
fish (Zabik and Zabik, 1996).
• For PCBs, residue reductions during cooking (baking and charbroiling) of the
homologues with the lowest and the highest numbers of chlorines (trichloro-,
tetrachloro- and octachloro-PCBs) tended to be less than residue reductions
for the pentachloro-, hexachloro- and heptachloro-PCBs, which typically
make up the major portion of the PCBs found in fish samples (Zabik and
Zabik, 1996).
• In general for heavy metals, tissue residues are not significantly reduced by
processing or cooking methods (Gutenmann and Lisk, 1991; Zabik and
Zabik, 1996).
The results of a number of fish preparation and cooking studies are presented in
Tables C-1 and C-2 for a variety of fish species. The data are relevant primarily
to concentrations in the standard fillet. Dose modification will depend on how the
dose is determined initially (i.e., what portion of the fish was analyzed to
determine contamination concentrations). Note that contaminants distributed
throughout the fish muscle tissue, such as mercury, will not be substantially
reduced through most fish preparation or cooking methods.
Table C-1 summarizes various study results where specific activities reduce
contaminants in standard fillets of fish species. Study citations are provided for
readers who wish to obtain more information on study methods and results.
Similar information obtained from studies of standard fillet, whole fish, or other
fillet types is presented in Table C-2. Both show that a high level of variability
should be expected in the effectiveness of skinning, trimming, and cooking fish.
The average reductions are reported for each study. Although significant
variability in percent reductions was found within each study, the mean reduction
data suggest that significant reductions can occur with food preparation and
cooking (Voilandetal., 1991). The cooked weight of fish tissue is always less than
the uncooked weight. On average, cooking reduces the fish weight by about one-
third (Great Lakes Sport Fish Advisory Task Force, 1993); therefore, the standard
C-9
-------
APPENDIX C
meal of 1/2 pound of raw fillet weighs about 1/3 pound after cooking. Most of the
weight reduction is due to water loss, but fat liquification and volatilization also
contribute to weight reduction (Great Lakes Sport Fish Advisory Task Force,
1993). The actual weight loss depends on the cooking technique used.
The results of studies shown in Tables C-1 through C-3 do not address chemical
degradation due to heat applied in cooking. Zabik et al. (1994) found that smoking
lake trout reduced pesticides and total PCBs significantly more than other cooking
methods, but this cooking method resulted in the formation of PAHs. Until there
is more information about the toxicity of the byproducts generated during the
degradation of PCBs, dioxins/furans, organochlorine pesticides, or the other
chemicals of concern, EPA recommends that no dose modification be assumed
due to degradation alone.
Zabik et al. (1994) found similarities in the percentage of pesticide and total PCB
reductions (ranging from 27.9 to 36.5 percent) attributed to cooking for Great
Lakes carp, salmon, lake trout, walleye, and white bass analyzed (Table C-3).
However, they assessed only lipophilic chlorinated hydrocarbons. Similarities in
their chemical behavior may be responsible for the similarities observed in the
study results listed in Table C-3. The information provided in this table is not
species-specific, which may limit the situations to which it is applicable.
C-10
-------
APPENDIX C
Table C-1. Summary of Contaminant Reductions Due to Skinning, Trimming, and
Cooking (Based on Standard Fillet)
Species
Brown Trout
Carp
Carp
(Great Lakes)
(Lake Erie)
Carp
(Lake Erie)
Contaminant
DDE
DDE
DDE
Mi rex
Mi rex
Mi rex
Mi rex
Mi rex
PCB
PCB
PCB
PCB
PCB
a-Chlordane
a-Chlordane
a-Chlordane
a-Chlordane
Dieldrin
Dieldrin
Dieldrin
Dieldrin
Heptachlor epoxide
PCB
PCB
PCB
TCDD
TCDD
p,p'-DDE
p,p'-DDE
p,p'-DDD
p,p'-DDD
a- chlordane
a- chlordane
y- chlordane
y- chlordane
cis-nonachlor
cis-nonachlor
trans-nonachlor
trans-nonachlor
HCB
HCB
dieldrin
dieldrin
Total PCBs
Activity3
Trimming
Smoking
Broiling
Trimming
Trimming
Smoking
Broiling
Trimming & cooking
Trimming
Trimming
Smoking
Broiling
Trimming & cooking
Skin-off & deep frying
Skin-off & pan frying
Skin-on & deep frying
Skin-on & pan frying
Skin-off & deep frying
Skin-off & pan frying
Skin-on & deep frying
Skin-on & pan frying
Skin-on & pan frying
Skin-off & deep frying
Skin-off & pan frying
Skin-on & deep frying
skin-on & cooked
skin off & cooked
skin-on & deep fried
skin off & deep fried
skin-on & deep fried
skin off & deep fried
skin-on & deep fried
skin off & deep fried
skin-on & deep fried
skin off & deep fried
skin-on & deep fried
skin off & deep fried
skin-on & deep fried
skin off & deep fried
skin-on & deep fried
skin off & deep fried
skin-on & deep fried
skin off & deep fried
skin-on & deep fried
Reduction
(%)b
52
27
20
44
45
39
26
74
46
43
27
0
78
44
17
38
51
76
58
56
59
82
37
25
38
approx. 37
approx. 54
28
45
30
35
37
56
32
41
34
53
54*
27
14
54
52
53
16
Reference
Skeaetal. (1979)
Skeaetal. (1979)
Skeaetal. (1979)
Voiland et al. (1991)
Skeaetal. (1979)
Skeaetal. (1979)
Skeaetal. (1979)
Skeaetal. (1979)
Voiland et al. (1991)
Skeaetal. (1979)
Skeaetal. (1979)
Skeaetal. (1979)
Skeaetal. (1979)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabik&Zabik1995
Zabik&Zabik1995
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
See footnotes at end of table.
(continued)
C-11
-------
APPENDIX C
Table C-1. (Continued)
Species Contaminant
Carp (con.) Total PCBs
(Lake Erie) p,p'-DDE
p,p'-DDE
p,p'-DDD
p,p'-DDD
a- chlordane
a- chlordane
y- chlordane
oxychlordane
cis-nonachlor
trans-nonachlor
trans-nonachlor
HCB
dieldrin
dieldrin
heptachlor epoxide
Total PCBs
Total PCBs
Carp p,p'-DDE
(Lake Huron) p,p'-DDE
p,p'-DDD
p,p'-DDD
a- chlordane
a- chlordane
y- chlordane
cis-nonachlor
trans-nonachlor
trans-nonachlor
HCB
HCB
dieldrin
Total PCBs
Total PCBs
p,p'-DDE
p,p'-DDE
p,p'-DDD
p,p'-DDD
a- chlordane
a- chlordane
y- chlordane
cis-nonachlor
trans-nonachlor
trans-nonachlor
HCB
HCB
dieldrin
Total PCBs
Activity3
skin off & deep fried
skin-on & pan fried
skin off & pan fried
skin-on & pan fried
skin off & pan fried
skin-on & pan fried
skin off & pan fried
skin-on & pan fried
skin-on & pan fried
skin-on & pan fried
skin-on & pan fried
skin off & pan fried
skin-on & pan fried
skin-on & pan fried
skin off & pan fried
skin-on & pan fried
skin-on & pan fried
skin off & pan fried
skin-on & deep fried
skin off & deep fried
skin-on & deep fried
skin off & deep fried
skin-on & deep fried
skin off & deep fried
skin-on & deep fried
skin-on & deep fried
skin-on & deep fried
skin off & deep fried
skin-on & deep fried
skin off & deep fried
skin-on & deep fried
skin-on & deep fried
skin off & deep fried
skin-on & pan fried
skin off & pan fried
skin-on & pan fried
skin off & pan fried
skin-on & pan fried
skin off & pan fried
skin-on & pan fried
skin-on & pan fried
skin-on & pan fried
skin off & pan fried
skin-on & pan fried
skin off & pan fried
skin-on & pan fried
skin-on & pan fried
Reduction
(%)b
32
36
17
54
40
43
26
20
38
42
7
3*
138*
27
19
8
22
19
46
39
31
51
32
33
29
54
13*
27
33
27
44
67
32
48
50
38
17
55
35
50
54
35
39
19
10*
93
42
Reference
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
See footnotes at end of table.
(continued)
C-12
-------
APPENDIX C
Table C-1. (Continued)
Species
Carp (con.)
(Lake Huron)
Chinook
Salmon
Chinook Salmon
(Great Lakes)
Chinook Salmon
(Lake Huron)
Contaminant
Total PCBs
PCB
a-Chlordane
a-Chlordane
a-Chlordane
a-Chlordane
a-Chlordane
a-Chlordane
a-Chlordane
Dieldrin
Dieldrin
Dieldrin
Dieldrin
Dieldrin
Dieldrin
Dieldrin
Heptachlor epoxide
Heptachlor epoxide
Heptachlor epoxide
Heptachlor epoxide
Heptachlor epoxide
Heptachlor epoxide
Heptachlor epoxide
PCB
PCB
PCB
PCB
PCB
PCB
PCB
Toxaphene
Toxaphene
Toxaphene
Toxaphene
Toxaphene
Toxaphene
Toxaphene
TCDD
TCDD
p,p'-DDT
p,p'-DDT
p,p'-DDE
p,p'-DDE
p,p'-DDD
p,p'-DDD
a- chlordane
a- chlordane
Activity3
skin off & pan fried
Skin-on & pan frying
Skin-off & baking
Skin-off & charbroiling
Skin-off & charbroiling after scoring
Skin-off & canning
Skin-on & baking
Skin-on & charbroiling
Skin-on & charbroiling after scoring
Skin-off & baking
Skin-off & charbroiling
Skin-off & charbroiling after scoring
Skin-off & canning
Skin-on & baking
Skin-on & charbroiling
Skin-on & charbroiling after scoring
Skin-off & baking
Skin-off & charbroiling
Skin-off & charbroiling after scoring
Skin-off & canning
Skin-on & baking
Skin-on & charbroiling
Skin-on & charbroiling after scoring
Skin-off & baking
Skin-off & charbroiling
Skin-off & charbroiling after scoring
Skin-off & canning
Skin-on & baking
Skin-on & charbroiling
Skin-on & charbroiling after scoring
Skin-off & baking
Skin-off & charbroiling
Skin-off & charbroiling after scoring
Skin-off & canning
Skin-on & baking
Skin-on & charbroiling
Skin-on & charbroiling after scoring
skin-on & cooked
skin off & cooked
skin-on & baked
skin-off & baked
skin-on & baked
skin-off & baked
skin-on & baked
skin-off & baked
skin-on & baked
skin-off & baked
Reduction
(%)b
37
31
44
41
45
37
27
42
51
30
31
40
40
29
40
50
52
40
42
37
23
45
48
38
44
46
36
33
40
49
34
30
34
74
22
37
47
approx. 43
approx. 57
23
26
35
47
27
4
33
51
Reference
Zabiketal. 1995b
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabik&Zabik1995
Zabik&Zabik1995
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
See footnotes at end of table.
(continued)
C-13
-------
APPENDIX C
Table C-1. (Continued)
Species Contaminant
Chinook Salmon y - chlordane
(Lake Huron) y- chlordane
(con.) oxychlordane
oxychlordane
cis-nonachlor
cis-nonachlor
trans-nonachlor
trans-nonachlor
HCB
HCB
dieldrin
dieldrin
heptachlor epoxide
heptachlor epoxide
toxaphene
toxaphene
Total PCBs
Total PCBs
p,p'-DDT
p,p'-DDT
p,p'-DDE
p,p'-DDE
p,p'-DDD
p,p'-DDD
a- chlordane
a- chlordane
y- chlordane
y- chlordane
oxychlordane
oxychlordane
cis-nonachlor
cis-nonachlor
trans-nonachlor
trans-nonachlor
HCB
HCB
dieldrin
dieldrin
heptachlor epoxide
heptachlor epoxide
toxaphene
toxaphene
Total PCBs
Total PCBs
p,p'-DDT
p,p'-DDT
p,p'-DDE
Activity3
skin-on & baked
skin-off & baked
skin-on & baked
skin-off & baked
skin-on & baked
skin-off & baked
skin-on & baked
skin-off & baked
skin-on & baked
skin-off & baked
skin-on & baked
skin-off & baked
skin-on & baked
skin-off & baked
skin-on & baked
skin-off & baked
skin-on & baked
skin-off & baked
skin-on & charbroiled
skin-off & charbroiled
skin-on & charbroiled
skin-off & charbroiled
skin-on & charbroiled
skin-off & charbroiled
skin-on & charbroiled
skin-off & charbroiled
skin-on & charbroiled
skin-off & charbroiled
skin-on & charbroiled
skin-off & charbroiled
skin-on & charbroiled
skin-off & charbroiled
skin-on & charbroiled
skin-off & charbroiled
skin-on & charbroiled
skin-off & charbroiled
skin-on & charbroiled
skin-off & charbroiled
skin-on & charbroiled
skin-off & charbroiled
skin-on & charbroiled
skin-off & charbroiled
skin-on & charbroiled
skin-off & charbroiled
skin-on, scored & charbroiled
skin-off , scored & charbroiled
skin-on, scored & charbroiled
Reduction
(%)b Reference
33
43
42
50
31
46
43
41
48
60
38
35
36
44
38
49
49
48
35
50
41
61
39
62
44
63
38
48
62
59
45
61
45
61
47
49
47
51
45
55
41
47
40
62
58
59
59
Zabiket al.
Zabik et al.
Zabiket al.
Zabiket al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabiket al.
Zabik et al.
Zabiket al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
See footnotes at end of table.
(continued)
C-14
-------
APPENDIX C
Table C-1. (Continued)
Species
Chinook
Salmon (con.)
(Lake Huron)
Chinook Salmon
(Lake Michigan)
Contaminant
p,p'-DDE
p,p'-DDD
p,p'-DDD
a- chlordane
a- chlordane
oxychlordane
oxychlordane
cis-nonachlor
cis-nonachlor
trans-nonachlor
trans-nonachlor
HCB
HCB
dieldrin
dieldrin
heptachlor epoxide
heptachlor epoxide
toxaphene
toxaphene
Total PCBs
Total PCBs
p,p'-DDT
p,p'-DDE
p,p'-DDD
a- chlordane
y- chlordane
oxychlordane
cis-nonachlor
trans-nonachlor
HCB
dieldrin
toxaphene
Total PCBs
p,p'-DDT
p,p'-DDT
p,p'-DDE
p,p'-DDE
p,p'-DDD
p,p'-DDD
a- chlordane
a- chlordane
y- chlordane
y- chlordane
oxychlordane
oxychlordane
cis-nonachlor
cis-nonachlor
Activity3
skin-off , scored & charbroiled
skin-on, scored & charbroiled
skin-off , scored & charbroiled
skin-on, scored & charbroiled
skin-off , scored & charbroiled
skin-on, scored & charbroiled
skin-off, scored & charbroiled
skin-on, scored & charbroiled
skin-off, scored & charbroiled
skin-on, scored & charbroiled
skin-off, scored & charbroiled
skin-on, scored & charbroiled
skin-off , scored & charbroiled
skin-on, scored & charbroiled
skin-off , scored & charbroiled
skin-on, scored & charbroiled
skin-off, scored & charbroiled
skin-on, scored & charbroiled
skin-off, scored & charbroiled
skin-on, scored & charbroiled
skin-off, scored & charbroiled
canned with skin-off
canned with skin-off
canned with skin-off
canned with skin-off
canned with skin-off
canned with skin-off
canned with skin-off
canned with skin-off
canned with skin-off
canned with skin-off
canned with skin-off
canned with skin-off
skin-on & baked
skin-off & baked
skin-on & baked
skin-off & baked
skin-on & baked
skin-off & baked
skin-on & baked
skin-off & baked
skin-on & baked
skin-off & baked
skin-on & baked
skin-off & baked
skin-on & baked
skin-off & baked
Reduction
(%)b Reference
51
54
57
63
54
49
50
49
50
63
54
50
46
62
43
57
56
56
48
61
52
80
38
8*
51
56
47
46
53
27
88
77
33
23
26
16
26
10
30
29
31
28
22
2810
2236
32
33
Zabiket al.
Zabiket al.
Zabiket al.
Zabiket al.
Zabik et al.
Zabiket al.
Zabiket al.
Zabiket al.
Zabiket al.
Zabiket al.
Zabiket al.
Zabiket al.
Zabiket al.
Zabiket al.
Zabiket al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabiket al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
See footnotes at end of table.
(continued)
C-15
-------
APPENDIX C
Table C-1. (Continued)
Reduction
Species
Chinook Salmon
(Lake Michigan)
(con.)
Chinook Salmon
(Lake Huron)
Chinook Salmon
(Lake Michigan)
Chinook Salmon
(Lake Huron)
Chinook Salmon
(Lake Michigan)
Contaminant
trans-nonachlor
trans-nonachlor
HCB
HCB
dieldrin
dieldrin
heptachlor epoxide
heptachlor epoxide
toxaphene
toxaphene
Total PCBs
Total PCBs
p,p'-DDT
p,p'-DDT
p,p'-DDE
p,p'-DDE
p,p'-DDD
p,p'-DDD
a- chlordane
a- chlordane
y- chlordane
y- chlordane
oxychlordane
oxychlordane
cis-nonachlor
cis-nonachlor
trans-nonachlor
trans-nonachlor
HCB
HCB
dieldrin
dieldrin
heptachlor epoxide
heptachlor epoxide
toxaphene
toxaphene
Total PCBs
Total PCBs
p,p'-DDT
p,p'-DDT
p,p'-DDE
p,p'-DDE
p,p'-DDD
p,p'-DDD
a- chlordane
Activity3
skin-on & baked
skin-off & baked
skin-on & baked
skin-off & baked
skin-on & baked
skin-off & baked
skin-on & baked
skin-off & baked
skin-on & baked
skin-off & baked
skin-on & baked
skin-off & baked
skin-on & charbroiled
skin-off & charbroiled
skin-on & charbroiled
skin-off & charbroiled
skin-on & charbroiled
skin-off & charbroiled
skin-on & charbroiled
skin-off & charbroiled
skin-on & charbroiled
skin-off & charbroiled
skin-on & charbroiled
skin-off & charbroiled
skin-on & charbroiled
skin-off & charbroiled
skin-on & charbroiled
skin-off & charbroiled
skin-on & charbroiled
skin-off & charbroiled
skin-on & charbroiled
skin-off & charbroiled
skin-on & charbroiled
skin-off & charbroiled
skin-on & charbroiled
skin-off & charbroiled
skin-on & charbroiled
skin-off & charbroiled
skin-on, scored, & charbroiled
skin-off , scored, & charbroiled
skin-on, scored, & charbroiled
skin-off , scored, & charbroiled
skin-on, scored, & charbroiled
skin-off , scored, & charbroiled
skin-on, scored, & charbroiled
(%)b Reference
28
28
34
27
21
25
14
32
7
22
25
29
48
23
41
30
48
20
43
27
43
29
46
21
49
31
43
21
53
40
39
12
48
29
33
16
44
33
54
45
35
34
34
42
46
Zabiket al.
Zabiket al.
Zabiket al.
Zabiket al.
Zabiket al.
Zabiket al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabiket al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
Zabik et al.
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
1995b
See footnotes at end of table.
(continued)
C-16
-------
APPENDIX C
Table C-1. (Continued)
Species
Chinook Salmon
(Lake Michigan)
(con.)
Chinook Salmon
(Lake Huron)
Chinook Salmon
(Lake Michigan)
Lake Trout
Contaminant
a- chlordane
y- chlordane
y- chlordane
oxychlordane
oxychlordane
cis-nonachlor
cis-nonachlor
trans-nonachlor
trans-nonachlor
HCB
HCB
dieldrin
dieldrin
heptachlor epoxide
heptachlor epoxide
toxaphene
toxaphene
Total PCBs
Total PCBs
p,p'-DDT
p,p'-DDE
p,p'-DDD
a- chlordane
y- chlordane
oxychlordane
cis-nonachlor
trans-nonachlor
HCB
dieldrin
heptachlor epoxide
toxaphene
Total PCBs
a-Chlordane
a-Chlordane
a-Chlordane
a-Chlordane
DDT
DDT
DDT
DDT
Dieldrin
Dieldrin
Dieldrin
Dieldrin
Heptachlor epoxide
Heptachlor epoxide
Activity3
skin-off , scored, & charbroiled
skin-on, scored, & charbroiled
skin-off , scored, & charbroiled
skin-on, scored, & charbroiled
skin-off , scored, & charbroiled
skin-on, scored, & charbroiled
skin-off , scored, & charbroiled
skin-on, scored, & charbroiled
skin-off , scored, & charbroiled
skin-on, scored, & charbroiled
skin-off , scored, & charbroiled
skin-on, scored, & charbroiled
skin-off , scored, & charbroiled
skin-on, scored, & charbroiled
skin-off , scored, & charbroiled
skin-on, scored, & charbroiled
skin-off , scored, & charbroiled
skin-on, scored, & charbroiled
skin-off , scored, & charbroiled
canned with skin-off
canned with skin-off
canned with skin-off
canned with skin-off
canned with skin-off
canned with skin-off
canned with skin-off
canned with skin-off
canned with skin-off
canned with skin-off
canned with skin-off
canned with skin-off
canned with skin-off
Skin-off & baking
Skin-off & charbroiling
Skin-off & salt boiling
Skin-on & smoking
Skin-off & baking
Skin-off & charbroiling
Skin-off & salt boiling
Skin-on & smoking
Skin-off & baking
Skin-off & charbroiling
Skin-off & salt boiling
Skin-on & smoking
Skin-off & baking
Skin-off & charbroiling
Reduction
(%)b
39
47
32
34
33
51
41
37
44
31
43
42
41
42
31
37
22
37
44
141 *
37
34*
35
35
30
28
43
33
43
28
72
39
26
41
6
53
14
21
1
60
8
15
16
43
39
39
Reference
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. 1995b
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
See footnotes at end of table.
(continued)
C-17
-------
APPENDIX C
Table C-1. (Continued)
Reduction
Species
Lake Trout (con.)
Lake Trout
(Great Lakes)
Lake Trout / Lean
(Lake Huron)
Lake Trout / Lean
(Lake Michigan)
Contaminant
Heptachlor epoxide
Heptachlor epoxide
PCB
PCB
PCB
PCB
Toxaphene
Toxaphene
Toxaphene
Toxaphene
TCDD
p,p'-DDT
p,p'-DDT
p,p'-DDE
p,p'-DDE
p,p'-DDD
p,p'-DDD
a- chlordane
a- chlordane
y- chlordane
y- chlordane
oxychlordane
oxychlordane
cis-nonachlor
cis-nonachlor
trans-nonachlor
trans-nonachlor
HCB
HCB
dieldrin
dieldrin
heptachlor epoxide
heptachlor epoxide
toxaphene
toxaphene
Total PCBs
Total PCBs
p,p'-DDT
p,p'-DDT
p,p'-DDE
p,p'-DDE
p,p'-DDD
p,p'-DDD
a- chlordane
a- chlordane
y- chlordane
Activity3
Skin-off & salt boiling
Skin-on & smoking
Skin-off & baking
Skin-off & charbroiling
Skin-off & salt boiling
Skin-on & smoking
Skin-off & baking
Skin-off & charbroiling
Skin-off & salt boiling
Skin-on & smoking
skin-off & cooked
skin-off & baked
skin-off & charbroiled
skin-off & baked
skin-off & charbroiled
skin-off & baked
skin-off & charbroiled
skin-off & baked
skin-off & charbroiled
skin-off & baked
skin-off & charbroiled
skin-off & baked
skin-off & charbroiled
skin-off & baked
skin-off & charbroiled
skin-off & baked
skin-off & charbroiled
skin-off & baked
skin-off & charbroiled
skin-off & baked
skin-off & charbroiled
skin-off & baked
skin-off & charbroiled
skin-off & baked
skin-off & charbroiled
skin-off & baked
skin-off & charbroiled
skin-off & baked
skin-off & charbroiled
skin-off & baked
skin-off & charbroiled
skin-off & baked
skin-off & charbroiled
skin-off & baked
skin-off & charbroiled
skin-off & baked
(%)b
3
59
13
29
10
46
31
40
5
51
61
17
34
18
9
6
16
7
18
83
38
6
12
17
18
19
16
15
23
8
30
4
12
18*
13
18
15
11
19
9
14
11
9
4
3
2
Reference
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabik&Zabik1995
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
See footnotes at end of table.
(continued)
C-18
-------
APPENDIX C
Table C-1. (Continued)
Species
Lake Trout / Lean
(Lake Michigan)
(con.)
Contaminant
y- chlordane
oxychlordane
oxychlordane
cis-nonachlor
cis-nonachlor
trans-nonachlor
trans-nonachlor
HCB
HCB
dieldrin
dieldrin
heptachlor epoxide
heptachlor epoxide
toxaphene
toxaphene
Total PCBs
Total PCBs
p,p'-DDT
p,p'-DDE
p,p'-DDD
a- chlordane
y- chlordane
oxychlordane
cis-nonachlor
trans-nonachlor
HCB
dieldrin
heptachlor epoxide
toxaphene
Total PCBs
p,p'-DDT
p,p'-DDE
p,p'-DDD
a- chlordane
y- chlordane
oxychlordane
cis-nonachlor
trans-nonachlor
HCB
dieldrin
heptachlor epoxide
toxaphene
Total PCBs
p,p'-DDT
p,p'-DDE
p,p'-DDD
a- chlordane
Activity3
skin-off & charbroiled
skin-off & baked
skin-off & charbroiled
skin-off & baked
skin-off & charbroiled
skin-off & baked
skin-off & charbroiled
skin-off & baked
skin-off & charbroiled
skin-off & baked
skin-off & charbroiled
skin-off & baked
skin-off & charbroiled
skin-off & baked
skin-off & charbroiled
skin-off & baked
skin-off & charbroiled
skin-off & baked
skin-off & baked
skin-off & baked
skin-off & baked
skin-off & baked
skin-off & baked
skin-off & baked
skin-off & baked
skin-off & baked
skin-off & baked
skin-off & baked
skin-off & baked
skin-off & baked
skin-off & charbroiled
skin-off & charbroiled
skin-off & charbroiled
skin-off & charbroiled
skin-off & charbroiled
skin-off & charbroiled
skin-off & charbroiled
skin-off & charbroiled
skin-off & charbroiled
skin-off & charbroiled
skin-off & charbroiled
skin-off & charbroiled
skin-off & charbroiled
skin-off & salt boiled
skin-off & salt boiled
skin-off & salt boiled
skin-off & salt boiled
Reduction
(%)b
3
11
11
18
10
2
9
19
15
18
7
12
5
13
15
10
7
12
9
11
4
3*
11
18
2
19
18
12
13
10
19
14
9
3
3
11
10
9
15
7
5
15
7
1 *
7
5
5
Reference
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
See footnotes at end of table.
(continued)
C-19
-------
APPENDIX C
Table C-1. (Continued)
Species
Lake Trout / Lean
(Lake Michigan)
(con.)
Lake Trout / Lean
(Lake Ontario)
Lake Trout
Siscowet/
High Fat Content
(Lake Superior)
Contaminant
y- chlordane
oxychlordane
cis-nonachlor
trans-nonachlor
HCB
dieldrin
heptachlor epoxide
toxaphene
Total PCBs
p,p'-DDT
p,p'-DDE
p,p'-DDD
a- chlordane
y- chlordane
oxychlordane
cis-nonachlor
trans-nonachlor
HCB
dieldrin
heptachlor epoxide
toxaphene
Fluoranthene
Total PCBs
Benzo[b]fluorene
3,6-
Dimethylphenanthrene
Benz[a]anthacene
Chrysene
Total PAHs
p,p'-DDT
p,p'-DDT
p,p'-DDE
p,p'-DDE
p,p'-DDD
p,p'-DDD
HCB
HCB
dieldrin
dieldrin
toxaphene
Total PCBs
Total PCBs
p,p'-DDT
p,p'-DDT
p,p'-DDE
p,p'-DDE
p,p'-DDD
Activity3
skin-off & salt boiled
skin-off & salt boiled
skin-off & salt boiled
skin-off & salt boiled
skin-off & salt boiled
skin-off & salt boiled
skin-off & salt boiled
skin-off & salt boiled
skin-off & salt boiled
skin-off & smoked
skin-off & smoked
skin-off & smoked
skin-off & smoked
skin-off & smoked
skin-off & smoked
skin-off & smoked
skin-off & smoked
skin-off & smoked
skin-off & smoked
skin-off & smoked
skin-off & smoked
skin-off & smoked
skin-off & smoked
skin-off & smoked
skin-off & smoked
skin-off & smoked
skin-off & smoked
skin-off & smoked
skin-off & baked
skin-off & charbroiled
skin-off & baked
skin-off & charbroiled
skin-off & baked
skin-off & charbroiled
skin-off & baked
skin-off & charbroiled
skin-off & baked
skin-off & charbroiled
skin-off & baked
skin-off & baked
skin-off & charbroiled
skin-off & baked
skin-off & charbroiled
skin-off & baked
skin-off & charbroiled
skin-off & baked
Reduction
(%)b
1
3
10
13
7
16
3
5
10
58
47
61
50
49
57
51
55
53
42
59
49
6782*
41
1170*
1245*
5582*
4086*
10058*
12
8
12
12
85
88
10
17
4
8
71
11
12
42
72
20
10
20
Reference
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
See footnotes at end of table.
(continued)
C-20
-------
APPENDIX C
Table C-1. (Continued)
Species
Lake Trout
Siscowet/
High Fat Content
(Lake Superior)
(con.)
Lake Trout
High Fat Content/
Siscowet
(Lake Superior)
(con.)
See footnotes at end
Contaminant
p,p'-DDD
a- chlordane
a- chlordane
y- chlordane
y- chlordane
oxychlordane
oxychlordane
cis-nonachlor
cis-nonachlor
trans-nonachlor
trans-nonachlor
HCB
HCB
dieldrin
dieldrin
heptachlor epoxide
heptachlor epoxide
toxaphene
toxaphene
Total PCBs
Total PCBs
p,p'-DDT
p,p'-DDE
p,p'-DDD
a- chlordane
y- chlordane
oxychlordane
cis-nonachlor
trans-nonachlor
HCB
dieldrin
heptachlor epoxide
toxaphene
Total PCBs
p,p'-DDT
p,p'-DDE
p,p'-DDD
a- chlordane
y- chlordane
oxychlordane
cis-nonachlor
trans-nonachlor
HCB
dieldrin
heptachlor epoxide
toxaphene
of table.
Activity3
skin-off & charbroiled
skin-off & baked
skin-off & charbroiled
skin-off & baked
skin-off & charbroiled
skin-off & baked
skin-off & charbroiled
skin-off & baked
skin-off & charbroiled
skin-off & baked
skin-off & charbroiled
skin-off & baked
skin-off & charbroiled
skin-off & baked
skin-off & charbroiled
skin-off & baked
skin-off & charbroiled
skin-off & baked
skin-off & charbroiled
skin-off & baked
skin-off & charbroiled
skin-off & baked
skin-off & baked
skin-off & baked
skin-off & baked
skin-off & baked
skin-off & baked
skin-off & baked
skin-off & baked
skin-off & baked
skin-off & baked
skin-off & baked
skin-off & baked
skin-off & baked
skin-off & charbroiled
skin-off & charbroiled
skin-off & charbroiled
skin-off & charbroiled
skin-off & charbroiled
skin-off & charbroiled
skin-off & charbroiled
skin-off & charbroiled
skin-off & charbroiled
skin-off & charbroiled
skin-off & charbroiled
skin-off & charbroiled
Reduction
(%)b
14
10
6
12
29*
9
18
17
21
9
18
16
24
15
16
57
3*
28
45
18
32
42
20
17
10
12
9
17
9
16
15
57
28
18
72
10
14
6
29*
18
21
18
24
16
3*
45
Reference
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
(continued)
C-21
-------
APPENDIX C
Table C-1. (Continued)
Species
Lake Trout
Siscowet/
High Fat Content
(Lake Superior)
(con.)
Smallmouth
Bass
Contaminant
Total PCBs
p,p'-DDT
p,p'-DDE
p,p'-DDD
a- chlordane
Y- chlordane
oxychlordane
cis-nonachlor
trans-nonachlor
HCB
dieldrin
heptachlor epoxide
toxaphene
Total PCBs
p,p'-DDT
p,p'-DDE
p,p'-DDD
a- chlordane
Y- chlordane
oxychlordane
cis-nonachlor
trans-nonachlor
HCB
dieldrin
heptachlor epoxide
toxaphene
Total PCBs
Phenathrene
Anthracene
Fluoranthene
Pyrene
Benzo[b]fluorene
3,6-
Dimethylphenanthrene
Benz[a]anthacene
Dibenz[ac]anthracene
Dibenzo[ae]pyrene
Dibenzo[ah]pyrene
Chrysene
Total PAHs
DDE
DDE
DDE
Mi rex
Mi rex
Mi rex
Mi rex
Activity3
skin-off & charbroiled
skin-off & salt boiled
skin-off & salt boiled
skin-off & salt boiled
skin-off & salt boiled
skin-off & salt boiled
skin-off & salt boiled
skin-off & salt boiled
skin-off & salt boiled
skin-off & salt boiled
skin-off & salt boiled
skin-off & salt boiled
skin-off & salt boiled
skin-off & salt boiled
skin-off & smoked
skin-off & smoked
skin-off & smoked
skin-off & smoked
skin-off & smoked
skin-off & smoked
skin-off & smoked
skin-off & smoked
skin-off & smoked
skin-off & smoked
skin-off & smoked
skin-off & smoked
skin-off & smoked
skin-off & smoked
skin-off & smoked
skin-off & smoked
skin-off & smoked
skin-off & smoked
skin-off & smoked
skin-off & smoked
skin-off & smoked
skin-off & smoked
skin-off & smoked
skin-off & smoked
skin-off & smoked
Trimming
Baking
Frying
Trimming
Baking
Frying
Trimming & cooking
Reduction
(%)b
32
16
25
18
24
28
22
11
13
38
12
10
17
19
61
42
44
43
40
63
45
45
46
41
35
44
37
10771 *
2677*
29654 *
5928*
255*
1260*
915*
259*
157*
8*
421 *
4173*
54
16
75
64
21
75
80
Reference
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Zabiketal. 1996
Skeaetal. (1979)
Skeaetal. (1979)
Skeaetal. (1979)
Skeaetal. (1979)
Skeaetal. (1979)
Skeaetal. (1979)
Skeaetal. (1979)
See footnotes at end of table.
(continued)
C-22
-------
APPENDIX C
Table C-1. (Continued)
Species
Smallmouth
Bass (con.)
White Bass
(Great Lakes)
Walleye
Walleye
(Great Lakes)
Walleye
(Lake Erie)
Walleye
(Lake Huron)
Walleye
(Lake Michigan)
See footnotes at end
Contaminant
PCB
PCB
PCB
PCB
TCDD
DDT
DDT
DDT
a-Chlordane
a-Chlordane
a-Chlordane
Dieldrin
Dieldrin
Dieldrin
PCB
PCB
PCB
Toxaphene
TCDD
Chlordane Complex
DDT Complex
Dieldrin
Total PCBs
Chlordane Complex
DDT Complex
Dieldrin
Total PCBs
Chlordane Complex
DDT Complex
Dieldrin
Total PCBs
Chlordane Complex
DDT Complex
Dieldrin
Total PCBs
Chlordane Complex
DDT Complex
Dieldrin
Total PCBs
Chlordane Complex
DDT Complex
Dieldrin
Total PCBs
Chlordane Complex
DDT Complex
of table.
Activity3
Trimming
Baking
Frying
Trimming & cooking
skin-on & cooked
Skin-on & baking
Skin-on & charbroiling
Skin-on & deep frying
Skin-on & baking
Skin-on & charbroiling
Skin-on & deep frying
Skin-on & baking
Skin-on & charbroiling
Skin-on & deep frying
Skin-on & baking
Skin-on & charbroiling
Skin-on & deep frying
Skin-on & baking
skin-on & cooked
skin-on & baked
skin-on & baked
skin-on & baked
skin-on & baked
skin-on & charbroiled
skin-on & charbroiled
skin-on & charbroiled
skin-on & charbroiled
skin-on & baked
skin-on & baked
skin-on & baked
skin-on & baked
skin-on & charbroiled
skin-on & charbroiled
skin-on & charbroiled
skin-on & charbroiled
skin-on & baked
skin-on & baked
skin-on & baked
skin-on & baked
skin-on & charbroiled
skin-on & charbroiled
skin-on & charbroiled
skin-on & charbroiled
skin-on & deep fat fried
skin-on & deep fat fried
Reduction
(%)b
64
16
74
80
approx. 80
4
16
11
32
33
-25
3
3
26
17
24
14
45
approx. 44
33
33
21
13
60
25
29
20
44
26
10
20
25
17
37
29
9
22
26
23
33
33
12
27
3
3
Reference
Skeaetal. (1979)
Skeaetal. (1979)
Skeaetal. (1979)
Skeaetal. (1979)
Zabik&Zabik1995
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabik&Zabik1995
Zabiketal. 1995a
Zabiketal. 1995a
Zabiketal. 1995a
Zabiketal. 1995a
Zabiketal. 1995a
Zabiketal. 1995a
Zabiketal. 1995a
Zabiketal. 1995a
Zabiketal. 1995a
Zabiketal. 1995a
Zabiketal. 1995a
Zabiketal. 1995a
Zabiketal. 1995a
Zabiketal. 1995a
Zabiketal. 1995a
Zabiketal. 1995a
Zabiketal. 1995a
Zabiketal. 1995a
Zabiketal. 1995a
Zabiketal. 1995a
Zabiketal. 1995a
Zabiketal. 1995a
Zabiketal. 1995a
Zabiketal. 1995a
Zabiketal. 1995a
Zabiketal. 1995a
(continued)
C-23
-------
APPENDIX C
Table C-1. (Continued)
Reduction
Species
Contaminant
Activity3
Reference
Walleye (con.)
(Lake Michigan)
Five species
(Great Lakes)
Dieldrin
Total PCBs
trichloro-PCB
trichloro-PCB
tetrachloro-PCB
tetrachloro-PCB
pentachloro-PCB
pentachloro-PCB
hexachloro-PCB
hexachloro-PCB
heptachloro-PCB
heptachloro-PCB
octachloro-PCB
octachloro-PCB
Total PCBs
Total PCBs
trichloro-PCB
trichloro-PCB
tetrachloro-PCB
tetrachloro-PCB
pentachloro-PCB
pentachloro-PCB
hexachloro-PCB
hexachloro-PCB
heptachloro-PCB
heptachloro-PCB
octachloro-PCB
octachloro-PCB
Total PCBs
Total PCBs
skin-on & deep fat fried
skin-on & deep fat fried
skin-on & baked
skin-off & baked
skin-on & baked
skin-off & baked
skin-on & baked
skin-off & baked
skin-on & baked
skin-off & baked
skin-on & baked
skin-off & baked
skin-on & baked
skin-off & baked
skin-on & baked
skin-off & baked
skin-on & charbroiled
skin-off & charbroiled
skin-on & charbroiled
skin-off & charbroiled
skin-on & charbroiled
skin-off & charbroiled
skin-on & charbroiled
skin-off & charbroiled
skin-on & charbroiled
skin-off & charbroiled
skin-on & charbroiled
skin-off & charbroiled
skin-on & charbroiled
skin-off & charbroiled
27
15
approx. 15
approx. 20
approx. 26
approx. 26.5
approx. 32
approx. 29
approx. 34
approx. 34.5
approx.
34.75
approx. 33
approx. 27
approx. 25
approx. 34
approx. 33
approx. 28
approx. 26
approx. 32
approx. 34
approx. 36
approx. 33
approx. 40
approx. 35
approx. 40
approx. 37
approx. 28
approx. 31
approx. 37
approx. 36
Zabiketal. 1995a
Zabiketal. 1995a
Zabik&Zabik1996
Zabik&Zabik1996
Zabik&Zabik1996
Zabik&Zabik1996
Zabik&Zabik1996
Zabik&Zabik1996
Zabik&Zabik1996
Zabik&Zabik1996
Zabik&Zabik1996
Zabik&Zabik1996
Zabik&Zabik1996
Zabik&Zabik1996
Zabik&Zabik1996
Zabik&Zabik1996
Zabik&Zabik1996
Zabik&Zabik1996
Zabik&Zabik1996
Zabik&Zabik1996
Zabik&Zabik1996
Zabik&Zabik1996
Zabik&Zabik1996
Zabik&Zabik1996
Zabik&Zabik1996
Zabik&Zabik1996
Zabik&Zabik1996
Zabik&Zabik1996
Zabik&Zabik1996
Zabik&Zabik1996
Skin-on refers to the trimming of only the belly flap; skin-off refers to the removal of the belly flap as well as the
lateral line and associated fat tissue.
Data from the Zabik et al. (1994) study were condensed by averaging contaminant reductions across lakes
whenever a fish species was sampled from more than one of the Great Lakes.
C-24
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APPENDIX C
Table C-2. Summary of Contaminant Reductions Due to Skinning, Trimming, and
Cooking (Based on Standard Fillet, Whole Fish or Other Fillet)
Species
American Shad
Bluefish
Chinook Salmon
Coho Salmon
Lake Trout
Perch
Winter Flounder
(Seafish)
Contaminant
DDT/DDE
PCB
PCB
PCB
PCB
PCB
PCB
PCB
Mi rex
PCB
PCB (1248)
PCB (1248)
PCB (1254)
PCB (1254)
DDT
DDT/DDE
DDT
Mi rex
PCB
PCB (1248)
PCB (1248)
PCB (1254)
PCB(1254)
Dieldrin
Dieldrin
DDT
DDT/DDE
DDT
DDT
DDT
DDT
DDT
DDT
Dieldrin
Mi rex
PCB
DDT
PCB
PCB
PCB
Activity Reduction (%]
Trimming
Trimming
Trimming
Baking
Broiling
Frying
Poaching
Trimming & cooking
Trimming
Trimming
Trimming & baking
Trimming & poaching
Trimming and baking
Trimming & poaching
Trimming
Trimming
Dressing
Trimming
Trimming
Trimming & baking
Trimming & poaching
Trimming & baking
Trimming & poaching
Roasted
Microwave
Trimming
Trimming
Dressing
Frying
Broiling
Broiling
Roasted
Microwave
Broiling
Trimming
Trimming
Dressing
Deep frying
Pan frying
Broiling
40
44
59
8
8
8
8
67
15
25
15
-1
-1
2
62
53
0
21
32
4
-9
-10
-14
25
47
54
46
0
64-72
64-72
39
30
54
48
50
50
90
47
-15
-17
|a Reference
NYSDEC(1981)
NYSDEC(1981)
Armbruster et al. (1989)°
Armbruster et al. (1989)°
Armbruster et al. (1989)°
Armbruster et al. (1989)°
Armbruster et al. (1989)°
Armbruster et al. (1989)°
NYSDEC(1981)
NYSDEC(1981)
Smith etal. (1973)
Smith etal. (1973)
Smith etal. (1973)
Smith etal. (1973)
Reinertetal. (1972)
NYSDEC(1981)
Reinertetal. (1972)
NYSDEC(1981)
NYSDEC(1981)
Smith etal. (1973)
Smith etal. (1973)
Smith etal. (1973)
Smith etal. (1973)
Zabiketal. (1994)
Zabiketal. (1994)
Reinertetal. (1972)
NYSDEC(1981)
Reinertetal. (1972)
Reinertetal. (1972)
Reinertetal. (1972)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
Zabiketal. (1994)
NYSDEC(1981)
NYSDEC(1981)
Reinertetal. (1972)
EPA (1992)
EPA (1992)
EPA (1992)
It could not be positively determined that reduction figures were calculated as changes in contaminant
concentrations from the standard fillet.
Average of findings reported in New York State Department of Environmental Conservation (1981) and White et
al. (1985).
Averages of findings reported in Armbruster et al. (1989).
C-25
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APPENDIX C
Table C-3. Average Contaminant Reductions Due to Cooking in Great
Lakes Fish a
Chemical Contaminant Reduction (%)
p,p'-DDT 34.0
p,p'-DDE 29.4
p,p'-DDD 29.0
a-Chlordane 34.8
y-Chlordane 33.0
Oxychlordane 35.6
c/s-Nonachlor 35.7
frans-Nonachlor 27.9
Dieldrin 28.7
Heptachlorepoxide 35.6
Toxaphene 36.5
Total PCBs 30.3
a Processing involved trimming the belly flap area for skin-on fillets and skinning and
removing fatty tissue from the belly flap area and the lateral line for skin-off fillets.
Source: Zabik et al. (1994).
C.3 REFERENCES
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trimming and cooking by several methods on polychlorinated biphenyls
(PCBs) residues in bluefish. J. Food Safety 9:235-244.
Barren, M.G., S.M. Plakas, and P.C. Wilga. 1991. Chlorpyrifos pharmacokinetics
and metabolism following intravascular and dietary administration in channel
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Branson, D.R., T.I. Takahashi, W.M. Parker, and G.E. Blau. 1985.
Bioconcentration kinetics of 2,3,7,8-tetrachlorodibenzo-p-dioxin in rainbow
trout. Environ. Toxicol. Chem. 4:779-788.
Bruggeman, W.A., A. Opperhuizen, A. Wijbenga, and O. Hutzinger. 1984.
Bioaccumulation of superlipophilic chemicals in fish. Toxicol. Environ. Chem.
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Gardner, A.M. and K.D. White. 1990. Polychlorinated dibenzofurans in the edible
portion of selected fish. Chemosphere 21 (1-2): 215-222.
C-26
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APPENDIX C
Great Lakes Sport Fish Advisory Task Force. 1993. Draft Protocol fora Uniform
Great Lakes Sport Fish Consumption Advisory. May.
Gutenmann, W.H., and D.J. Lisk. 1991. Higher average mercury concentration in
fish fillets after skinning and fat removal. J. Food Safety 11(2):99-103.
Gutenmann, W.H., J.G. Ebel Jr., H.T. Kuntz , K.S. Yourstone, and D.J. Lisk. I.
1992. Residues of p,p' DDE and mercury in lake trout as a function of age.
Arch. Environ. Contam. Toxicol. 22:452-455.
Harrison, S.E., and J.F. Klaverkamp. 1990. Metal contamination in liver and
muscle of northern pike (Esox lucius) and white sucker (Catostomus
commersoni) from lakes near the smelter at Flin Flon, Manitoba Canada .
Environ. Toxicol. Chem. 9:941-956.
Kleeman, J., J.R. Olson, S.S. Chen, and R.E. Peterson. 1986a. Metabolism and
disposition of 2,3,7,8-tetrachlorodibenzo-p-dioxin in rainbow trout. Toxicol.
Appl. Pharmacol. 83:391-401.
Kleeman, J., J.R. Olson, S.S. Chen, and R.E. Peterson. 1986b. 2,3,7,8-
tetrachlorodibenzo-p-dioxin metabolism and disposition in yellow perch.
Toxicol. Appl. Pharmacol. 83:402-411.
Lake, J.L, R. McKinney, C.A. Lake, F.A. Osterman, and J. Heltshe. 1995.
Comparison of patterns of polychlorinated biphenyl congeners in water,
sediment, and indigenous organisms from New Bedford Harbor,
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Marcovecchio, J.E., V.J. Moreno, and A. Perez. 1988. The sole, Paralichthyssp.,
as an indicator species for heavy metal pollution in the Bahia Blanca Estuary,
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Metcalf, T.L. and C.D. Metcalf. 1997. The trophodynamics of PCBs, including
mono- and non-ortho congeners, in the food web of North-Central Lake
Ontario. Sci. Total Environ. 201: 245-272.
Minnesota Department of Health. 1992. Minnesota Fish Consumption Advisory.
Minneapolis, MN. May.
Muir, D.C.G., A.L. Yarechewski, A. Knoll, and G.R.B. Webster. 1986.
Bioconcentration and disposition of 1,3,6,8-tetrachlorodibenzo-p-dioxin and
octachlorodibenzo-p-dioxin by rainbow trout and fathead minnows. Environ.
Toxicol. Chem. 5: 261-272.
Niimi, A.J., and B.G. Oliver. 1989. Distribution of polychlorinated biphenyl
congeners and other halocarbons in whole fish and muscle among Lake
Ontario salmonids. Environ. Sci. Technol. 23: 83-88.
C-27
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APPENDIX C
Norey, C.G., M.W. Brown, A. Cryer, and J. Kay. 1990. A Comparison of the
Accumulation, Tissue Distribution, and Secretion of Cadmium in Different
Species of Freshwater Fish. Compar. Biochem. Physiol. 96C(1): 181-184.
NYSDEC (New York State Department of Environmental Conservation). 1981.
Toxic Substances in Fish and Wildlife. Technical Report 81-1 (BEP). Division
of Fish and Wildlife, Albany, NY.
Oliver, E.G. and A.J. Niimi. 1988. Trophodynamic analysis of polychlorinated
biphenyl congeners and other chlorinated hydrocarbons in the Lake Ontario
Ecosystem. Environ. Sci. Technol. 22: 388-397.
Reinert, R.E., D. Stewart, and H.L. Seagran. 1972. Effects of dressing and
cooking on DDT concentrations in certain fish from Lake Michigan. JFish Res
Board Can 29:525-529.
Sanders, M. and B.L. Hayes. 1988. Distribution pattern and reduction of
polychlorinated biphenyls (PCB) in bluefish (Pomatomus saltatrix lineaus)
fillets through adipose tissue removal. Bull. Environ. Contam. Toxicol. 41:
670-677.
Skea, J.C., H.A. Simonin, E.J. Harris, S. Jackling, and J.J. Spagnoli. 1979.
Reducing levels of mirex, arochlor 1254, and DDE by trimming and cooking
Lake Ontario brown trout (Salmo trutta L.) and smallmouth bass (Micropterus
dolomieui lacepede). J Great Lakes Res. 5(2): 153-159.
Smith, W.E., K. Funk, and M.E. Zabik. 1973. Effects of cooking on concentrations
of PCB and DDT compounds in Chinook (Oncorhychus tshawytscha) and
coho (O. kisutch) salmon from Lake Michigan. J. Fish Res. Board Can. 30:
702-706.
Stern, G., G. Muir, C. Ford, N. Grift, E. Dewally, T. Bidleman, and M. Walls. 1992.
Isolation and identification of two major recalcitrant toxaphene congeners in
aquatic biota. Environ. Sci. Technol. 26:1838-1840.
U.S. EPA (Environmental Protection Agency). 1992. National Study of Chemical
Residues in Fish, Volumes I. EPA 823-R-92-008a. Office of Science and
Technology, Washington, DC. September.
U.S. EPA (Environmental Protection Agency). 1995a. Guidance for Assessing
Chemical Contamination Data for Use in Fish Advisories, Volume 1: Fish
Sampling and Analysis. Second Edition. EPA823-R-95-007. Off ice of Science
and Technology, Washington, DC. September.
C-28
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APPENDIX C
U.S. EPA (Environmental Protection Agency). 1995b. Guidance for Assessing
Chemical Contamination Data for Use in Fish Advisories, Volume 4: Risk
Communication. EPA-823-R-95-001. Office of Science and Technology,
Washington, DC. March.
Voiland Jr., M.P., K.L Gall, D.J. Lisk, and D.B. MacNeill 1991. Effectiveness of
recommended fat trimming procedures on the reduction of PCB and mirex
levels in brown trout (Salmo trutta) from Lake Ontario. J Great Lakes Res
17(4):454-460.
White, R.J, H.T. Kim, and J.S. Kim 1985. Polychlorinated biphenyls in striped
bass (Morone saxatilis) collected from the Hudson River, New York, U.S.A..
during fall 1981. Bull. Environ. Contam. Toxicol. 34:883-889.
Zabik, M.E., M.J. Zabik, and H. Humphrey. 1994. Assessment of Contaminants
in Five Species of Great Lakes Fish at the Dinner Table. Final Report to the
Great Lakes Protection Fund, Chicago, Illinois. March.
Zabik, M.E. and M.J. Zabik. 1995. Tetrachlorodibenzo-p-dioxin residue reduction
by cooking/processing of fish fillets harvested from the Great Lakes. Bull.
Environ. Contam. Toxicol. 55:264-269.
Zabik, M.E. and M.J. Zabik. 1996. Influence of processing on environmental
contaminants in foods. Food Technol. 50: 225-229.
Zabik, M.E., M.J.. Zabik, A.M. Booren, S. Daubenmire, M.A. Pascall, R. Welch,
and H. Humphrey. 1995a. Pesticides and total polychlorinated biphenyls
residues in raw and cooked walleye and white bass harvested from the Great
Lakes. Bull. Environ. Contam. Toxicol. 54: 396-402.
Zabik, M.E., M.J. Zabik, A.M. Booren, M. Nettles, J.H. Song, R. Welch and H.
Humphrey. 1995b. Pesticides and total polychlorinated biphenyls in Chinook
salmon and carp harvested from the Great Lakes: Effects of skin-on and skin-
off processing and selected cooking methods. J. Agric. Food Chem. 43:993-
1001.
Zabik, M.E., A.M. Booren, M.J. Zabik, R. Welch, and H. Humphrey. 1996.
Pesticide residues, PCBs and PAHs in baked, charbroiled, salt boiled, and
smoked Great Lakes lake trout. Food Chem. 55 (3): 231-239.
C-29
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APPENDIX D
GUIDANCE FOR RISK CHARACTERIZATION
-------
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UNITED STATES ENVIRONMENTAL PROTECTION AGENCY
WASHINGTON, D.C. 20460
THE ADMINISTRATOR
MAR 2 11995
MEMORANDUM
SUBJECT: EPA Risk Characterization Program
TO Assistant Administrators
Associate Administrators
Regional Administrators
General Counsel
Inspector General
EPA has achieved significant pollution reduction over the past 20 years, but the
challenges we face now are very different from those of the past. Many more people are aware of
environmental issues today than in the past and their level of sophistication and interest in
understanding these issues continues to increase. We now work with a populace which is not
only interested in knowing what EPA thinks about a particular issue, but also how we come to
our conclusions.
More and more key stakeholders in environmental issues want enough information to
allow them to independently assess and make judgments about the significance of environmental
risks and the reasonableness of our risk reduction actions. If we are to succeed and build our
credibility and stature as a leader in environmental protection for the next century, EPA must be
responsive and resolve to more openly and fully communicate to the public the complexities and
challenges of environmental decisionmaking in the face of scientific uncertainty.
As the issues we face become more complex, people both inside and outside of EPA must
better understand the basis for our decisions, as well as our confidence in the data, the science
policy judgments we have made, and the uncertainty in the information base. In order to achieve
this better understanding, we must improve the way in which we characterize and communicate
environmental risk. We must embrace certain fundamental values so that we may begin the
process of changing the way in which we interact with each other, the public, and key
stakeholders on environmental risk issues. I need your help to ensure that these values are
embraced and that we change the way we do business.
-------
-2-
First, we must adopt as values transparency in our decisionmaking process and clarity in
communication with each other and the public regarding environmental risk and the uncertainties
associated with our assessments of environmental risk. This means that we must fully, openly,
and clearly characterize risks. In doing so, we will disclose the scientific analyses, uncertainties,
assumptions, and science policies which underlie our decisions as they are made throughout the
risk assessment and risk management processes. I want to be sure that key science policy issues
are identified as such during the risk assessment process, that policy makers are fully aware and
engaged in the selection of science policy options, and that their choices and the rationale for
those choices are clearly articulated and visible in our communications about environmental risk.
I understand that some may be concerned about additional challenges and disputes. I
expect that we will see more challenges, particularly at first. However, I strongly believe that
making this change to a more open decisionmaking process will lead to more meaningful public
participation, better information for decisionmaking, improved decisions, and more public
support and respect for EPA positions and decisions. There is value in sharing with others the
complexities and challenges we face in making decisions in the face of uncertainty. I view
making this change as essential to the long-term success of this Agency.
Clarity in communication also means that we will strive to help the public put
environmental risk in the proper perspective when we take risk management actions. We must
meet this challenge and find legitimate ways to help the public better comprehend the relative
significance of environmental risks.
Second, because transparency in decisionmaking and clarity in communication will likely
lead to more outside questioning of our assumptions and science policies, we must be more
vigilant about ensuring that our core assumptions and science policies are consistent and
comparable across programs, well grounded in science, and that they fall within a "zone of
reasonableness."
While I believe that the American public expects us to err on the side of protection in the
face of scientific uncertainty, I do not want our assessments to be unrealistically conservative.
We cannot lead the fight for environmental protection into the next century unless we use
common sense in all we do.
These core values of transparency, clarity, consistency, and reasonableness need to guide
each of us in our day-to-day work; from the toxicologist reviewing the individual cancer study, to
the exposure and risk assessors, to the risk manager, and through to the ultimate decisionmaker. I
recognize that issuing this memo will not by itself result in any change. You need to believe in
the importance of this change and convey your beliefs to your managers and staff through your
words and actions in order for the change to occur. You also need to play an integral role in
developing the implementing policies and procedures for your programs.
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-3-
I am issuing the attached EPA Risk Characterization Policy and Guidance today. I view
these documents as building blocks for the development of your program-specific policies and
procedures. The Science Policy Council (SPC) plans to adopt the same basic approach to
implementation as was used for Peer Review. That is, the Council will form an Advisory Group
that will work with a broad Implementation Team made up of representatives from every
Program Office and Region. Each Program Office and each Region will be asked by the
Advisory Group to develop program and region-specific policies and procedures for risk
characterization consistent with the values of transparency, clarity, consistency, and
reasonableness and consistent with the attached policy and guidance.
I recognize that as you develop your Program-specific policies and procedures you are
likely to need additional tools to fully implement this policy. I want you to identify these needed
tools and work cooperatively with the Science Policy Council in their development. I want your
draft program and region-specific policies, procedures, and implementation plans to be
developed and submitted to the Advisory Group for review by no later than May 30, 1995. You
will be contacted shortly by the SPC Steering Committee to obtain the names of your nominees
to the Implementation Team.
CarorM. Browner
Attachments
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March 1995
POLICY FOR RISK CHARACTERIZATION
at the U.S. Environmental Protection Agency
INTRODUCTION
Many EPA policy decisions are based in part on the results of risk assessment, an analysis of
scientific information on existing and projected risks to human health and the environment. As
practiced at EPA, risk assessment makes use of many different kinds of scientific concepts and
data (e.g., exposure, toxicity, epidemiology, ecology), all of which are used to "characterize" the
expected risk associated with a particular agent or action in a particular environmental context.
Informed use of reliable scientific information from many different sources is a central feature of
the risk assessment process.
Reliable information may or may not be available for many aspects of a risk assessment.
Scientific uncertainty is a fact of life for the risk assessment process, and agency managers
almost always must make decisions using assessments that are not as definitive in all important
areas as would be desirable. They therefore need to understand the strengths and the limitations
of each assessment, and to communicate this information to all participants and the public.
This policy reaffirms the principles and guidance found in the Agency's 1992 policy (Guidance
on Risk Characterization for Risk Managers and Risk Assessors, February 26, 1992). That
guidance was based on EPA's risk assessment guidelines, which are products of peer review and
public comment. The 1994 National Research Council (NRC) report, "Science and Judgment in
Risk Assessment," addressed the Agency's approach to risk assessment, including the 1992 risk
characterization policy. The NRC statement accompanying the report stated, "... EPA's overall
approach to assessing risks is fundamentally sound despite often-heard criticisms, but the Agency
must more clearly establish the scientific and policy basis for risk estimates and better describe
the uncertainties in its estimates of risk."
This policy statement and associated guidance for risk characterization is designed to ensure that
critical information from each stage of a risk assessment is used in forming conclusions about
risk and that this information is communicated from risk assessors to risk managers (policy
makers), from middle to upper management, and from the Agency to the public. Additionally, the
policy will provide a basis for greater clarity, transparency, reasonableness, and consistency in
risk assessments across Agency programs. While most of the discussion and examples in this
policy are drawn from health risk assessment, these values also apply to ecological risk
assessment. A parallel effort by the Risk Assessment Forum to develop EPA ecological risk
assessment guidelines will include guidance specific to ecological risk characterization.
Policy Statement
Each risk assessment prepared in support of decision-making at EPA should include a
risk characterization that follows the principles and reflects the values outlined in this policy. A
-------
risk characterization should be prepared in a manner that is clear, transparent, reasonable and
consistent with other risk characterizations of similar scope prepared across programs in the
Agency. Further, discussion of risk in all EPA reports, presentations, decision packages, and
other documents should be substantively consistent with the risk characterization. The nature of
the risk characterization will depend upon the information available, the regulatory application
of the risk information, and the resources (including time) available. In all cases, however, the
assessment should identify and discuss all the major issues associated with determining the
nature and extent of the risk and provide commentary on any constraints limiting fuller
exposition.
Key Aspects of Risk Characterization
Bridging risk assessment and risk management. As the interface between risk
assessment and risk management, risk characterizations should be clearly presented, and
separate from any risk management considerations. Risk management options should be
developed using the risk characterization and should be based on consideration of all relevant
factors, scientific and nonscientific.
Discussing confidence and uncertainties. Key scientific concepts, data and methods
(e.g., use of animal or human data for extrapolating from high to low doses, use of
pharmacokinetics data, exposure pathways, sampling methods, availability of chemical-specific
information, quality of data) should be discussed. To ensure transparency, risk characterizations
should include a statement of confidence in the assessment that identifies all major
uncertainties along with comment on their influence on the assessment, consistent with the
Guidance on Risk Characterization (attached).
Presenting several types of risk information. Information should be presented on the
range of exposures derived from exposure scenarios and on the use of multiple risk descriptors
(e.,g., central tendency, high end of individual risk, population risk, important subgroups, if
known) consistent with terminology in the Guidance on Risk Characterization, Agency risk
assessment guidelines, and program-specific guidance. In decision-making, risk managers
should use risk information appropriate to their program legislation.
EPA conducts many types of risk assessments, including screening-level assessments of
new chemicals, in-depth assessments of pollutants such as dioxin and environmental tobacco
smoke, and site-specific assessments for hazardous waste sites. An iterative approach to risk
assessment, beginning with screening techniques, may be used to determine if a more
comprehensive assessment is necessary. The degree to which confidence and uncertainty are
addressed in a risk characterization depends largely on the scope of the assessment. In general,
the scope of the risk characterization should reflect the information presented in the risk
assessment and program-specific guidance. When special circumstances (e.g., lack of data,
extremely complex situations, resource limitations, statutory deadlines) preclude a full
-------
assessment, such circumstances should be explained and their impact on the risk assessment
discussed.
Risk Characterization in Context
Risk assessment is based on a series of questions that the assessor asks about scientific
information that is relevant to human and/or environmental risk. Each question calls for
analysis and interpretation of the available studies, selection of the concepts and data that are
most scientifically reliable and most relevant to the problem at hand, and scientific conclusions
regarding the question presented. For example health risk assessments involve the following
questions:
Hazard Identification—What is known about the capacity of an environmental
agent for causing cancer or other adverse health effects in humans, laboratory
animals, or wildlife species? What are the related uncertainties and science
policy choices?
Dose-Response Assessment—What is known about the biological mechanisms
and dose-response relationships underlying any effects observed in the
laboratory or epidemiology studies providing data for the assessment? What are
the related uncertainties and science policy choices?
Exposure Assessment—What is known about the principal paths, patterns, and
magnitudes of human or wildlife exposure and numbers of persons or wildlife
species likely to be exposed? What are the related uncertainties and science
policy choices?
Corresponding principles and questions for ecological risk assessment are being discussed as
part of the effort to develop ecological risk guidelines.
Risk characterization is the summarizing step of risk assessment. The risk
characterization integrates information from the preceding components of the risk assessment
and synthesizes an overall conclusion about risk that is complete, informative and useful for
decisionmakers.
Risk characterizations should clearly highlight both the confidence and the uncertainty
associated with the risk assessment. For example, numerical risk estimates should always be
accompanied by descriptive information carefully selected to ensure an objective and balanced
characterization of risk in risk assessment reports and regulatory documents. In essence, a risk
characterization conveys the assessor's judgment as to the nature and existence of (or lack of)
human health or ecological risks. Even though a risk characterization describes limitations in an
assessment, a balanced discussion of reasonable conclusions and related uncertainties enhances,
rather than detracts, from the overall credibility of each assessment.
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"Risk characterization" is not synonymous with "risk communication." This risk
characterization policy addresses the interface between risk assessment and risk management.
Risk communication, in contrast, emphasizes the process of exchanging information and
opinion with the public—including individuals, groups, and other institutions. The development
of a risk assessment may involve risk communication. For example, in the case of site-specific
assessments for hazardous waste sites, discussions with the public may influence the exposure
pathways included in the risk assessment. While the final risk assessment document (including
the risk characterization) is available to the public, the risk communication process may be
better served by separate risk information documents designed for particular audiences.
Promoting Clarity, Comparability and Consistency
There are several reasons that the Agency should strive for greater clarity, consistency
and comparability in risk assessments. One reason is to minimize confusion. For example,
many people have not understood that a risk estimate of one in a million for an "average"
individual is not comparable to another one in a million risk estimate for the "most exposed
individual." Use of such apparently similar estimates without further explanation leads to
misunderstandings about the relative significance of risks and the protectiveness of risk
reduction actions.
EPA's Exposure Assessment Guidelines provide standard descriptors of exposure and
risk. Use of these terms in all Agency risk assessments will promote consistency and
comparability. Use of several descriptors, rather than a single descriptor, will enable EPA to
present a fuller picture of risk that corresponds to the range of different exposure conditions
encountered by various individuals and populations exposed to most environmental chemicals.
Legal Effect
This policy statement and associated guidance on risk characterization do not establish
or affect legal rights or obligations. Rather, they confirm the importance of risk characterization
as a component of risk assessment, outline relevant principles, and identify factors Agency staff
should consider in implementing the policy.
The policy and associated guidance do not stand alone; nor do they establish a binding
norm that is finally determinative of the issues addressed. Except where otherwise provided by
law, the Agency's decision on conducting a risk assessment in any particular case is within the
Agency's discretion. Variations in the application of the policy and associated guidance,
therefore, are not a legitimate basis for delaying or complicating action on Agency decisions.
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Applicability
Except where otherwise provided by law and subject to the limitations on the policy's
legal effect discussed above, this policy applies to risk assessments prepared by EPA and to risk
assessments prepared by others that are used in support of EPA decisions.
EPA will consider the principles in this policy in evaluating assessments submitted to
EPA to complement or challenge Agency assessments. Adherence to this Agency-wide policy
will improve understanding of Agency risk assessments, lead to more informed decisions, and
heighten the credibility of both assessments and decisions.
Implementation
Assistant Administrators and Regional Administrators are responsible for
implementation of this policy within their organizational units. The Science Policy Council
(SPC) is organizing Agency-wide implementation activities. Its responsibilities include
promoting consistent interpretation, assessing Agency-wide progress, working with external
groups on risk characterization issues and methods, and developing recommendations for
revisions of the policy and guidance, as necessary.
Each Program and Regional office will develop office-specific policies and procedures
for risk characterization that are consistent with this policy and the associated guidance. Each
Program and Regional office will designate a risk manager or risk assessor as the office
representative to the Agency-wide Implementation Team, which will coordinate development
of office-specific policies and procedures and other implementation activities. The SPC will
also designate a small cross-Agency Advisory Group that will serve as the liaison between the
SPC and the Implementation Team.
In ensuring coordination and consistency among EPA offices, the Implementation Team
will take into account statutory and court deadlines, resource implications, and existing Agency
and program-specific guidance on risk assessment. The group will work closely with staff
throughout Headquarters and Regional offices to promote development of risk characterizations
that present a full and complete picture of risk that meets the needs of the risk managers.
MAR 2 1 1995
APPROVED: fL£4&K-* CSS.\/J
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ELEMENTS TO CONSIDER WHEN DRAFTING EPA RISK
CHARACTERIZATIONS
March 1995
Background—Risk Characterization Principles
There are a number of principles which form the basis for a risk characterization:
• Risk assessments should be transparent, in that the conclusions drawn from the science are
identified separately from policy judgements, and the use of default values or methods and
the use of assumptions in the risk assessment are clearly articulated.
• Risk characterizations should include a summary of the key issues and conclusions of each
of the other components of the risk assessment, as well as describe the likelihood of harm.
The summary should include a description of the overall strengths and the limitations
(including uncertainties) of the assessment and conclusions.
• Risk characterizations should be consistent in general format, but recognize the unique
characteristics of each specific situation.
• Risk characterizations should include, at least in a qualitative sense, a discussion of how a
specific risk and its context compares with other similar risks. This may be accomplished by
comparisons with other chemicals or situations in which the Agency has decided to act, or
with other situations which the public may be familiar with. The discussion should highlight
the limitations of such comparisons.
• Risk characterization is a key component of risk communication, which is an interactive
process involving exchange of information and expert opinion among individuals, groups
and institutions.
Conceptual Guide for Developing Chemical-Specific Risk Characterizations
The following outline is a guide and formatting aid for developing risk characterizations for
chemical risk assessments. Similar outlines will be developed for other types of risk
characterizations, including site-specific assessments and ecological risk assessments. A
common format will assist risk managers in evaluating and using risk characterization.
The outline has two parts. The first part tracks the risk assessment to bring forward its major
conclusions. The second part draws all of the information together to characterize risk. The
outline represents the expected findings for a typical complete chemical assessment for a single
chemical. However, exceptions for the circumstances of individual assessments exist and
should be explained as part of the risk characterization. For example, particular statutory
requirements, court-ordered deadlines, resource limitations, and other specific factors may be
described to explain why certain elements are incomplete.
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This outline does not establish or affect legal rights or obligations. Rather, it confirms the
importance of risk characterization, outlines relevant principles, and identifies factors Agency
staff should consider in implementing the policy. On a continuing basis, Agency management is
expected to evaluate the policy as well as the results of its application throughout the Agency
and undertake revisions as necessary. Therefore, the policy does not stand alone; nor does it
establish a binding norm that is finally determinative of the issues addressed. Minor variations
in its application from one instance to another are appropriate and expected; they thus are not a
legitimate basis for delaying or complicating action on otherwise satisfactory scientific,
technical, and regulatory products.
PART ONE
SUMMARIZING MAJOR CONCLUSIONS IN RISK CHARACTERIZATION
I. Characterization of Hazard Identification
A. What is the key toxicological study (or studies) that provides the basis for health
concerns?
- How good is the key study?
- Are the data from laboratory or field studies? In single species or multiple
species?
- If the hazard is carcinogenic, comment on issues such as: observation of single or
multiple tumor sites; occurrence of benign or malignant tumors; certain tumor
types not linked to carcinogen!city; use of the maximum tolerated dose (MTD).
- If the hazard is other than carcinogenic, what endpoints were observed, and what
is the basis for the critical effect?
- Describe other studies that support this finding.
- Discuss any valid studies which conflict with this finding.
B. Besides the health effect observed in the key study, are there other health endpoints
of concern?
- What are the significant data gaps?
C. Discuss available epidemiological or clinical data. For epidemiological studies:
- What types of studies were used, i.e., ecologic, case-control, cohort?
- Describe the degree to which exposures were adequately described.
- Describe the degree to which confounding factors were adequately accounted for.
- Describe the degree to which other causal factors were excluded.
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D. How much is known about how (through what biological mechanism) the chemical
produces adverse effects?
- Discuss relevant studies of mechanisms of action or metabolism.
- Does this information aid in the interpretation of the toxicity data?
- What are the implications for potential health effects?
E. Comment on any non-positive data in animals or people, and whether these data were
considered in the hazard identification.
F. If adverse health effects have been observed in wildlife species, characterize such
effects by discussing the relevant issues as in A through E above.
G. Summarize the hazard identification and discuss the significance of each of the
following:
- confidence in conclusions;
- alternative conclusions that are also supported by the data;
- significant data gaps; and
- highlights of major assumptions.
II. Characterization of Dose-Response
A. What data were used to develop the dose-response curve? Would the result have been
significantly different if based on a different data set?
- If animal data were used;
- which species were used? most sensitive, average of all species, or other?
- were any studies excluded? why?
- If epidemiological data were used:
- Which studies were used? only positive studies, all studies, or some other
combination?
- Were any studies excluded? why?
- Was a meta-analysis performed to combine the epidemiological studies? what
approach was used? were studies excluded? why?
B. What model was used to develop the dose-response curve? What rationale supports
this choice? Is chemical-specific information available to support this approach?
- For non-carcinogenic hazards:
- How was the RfD/RfC (or the acceptable range) calculated?
- What assumptions or uncertainty factors were used?
- What is the confidence in the estimates?
- For carcinogenic hazards:
- What dose-response model was used? LMS or other linear-at-low dose model,
a biologically based model based on metabolism data, or data about possible
mechanisms of action?
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- What is the basis for the selection of the particular dose-response model used?
Are there other models that could have been used with equal plausibility and
scientific validity? What is the basis for selection of the model used in this
instance?
C. Discuss the route and level of exposure observed, as compared to expected human
exposures.
- Are the available data from the same route of exposure as the expected human
exposures? If not, are pharmacokinetic data available to extrapolate across route
of exposure?
- How far does one need to extrapolate from the observed data to environmental
exposures (one to two orders of magnitude? multiple orders of magnitude)? What
is the impact of such an extrapolation?
D. If adverse health effects have been observed in wildlife species, characterize dose
response information using the process outlined in A-C.
III. Characterization of Exposure
A. What are the most significant sources of environmental exposure?
- Are there data on sources of exposure from different media? What is the relative
contribution of different sources of exposure?
- What are the most significant environmental pathways for exposure?
B. Describe the populations that were assessed, including as the general population,
highly exposed groups, and highly susceptible groups.
C. Describe the basis for the exposure assessment, including any monitoring, modeling,
or other analyses of exposure distributions such as Monte-Carlo or krieging.
D. What are the key descriptors of exposure?
- Describe the (range of) exposures to: "average" individuals, "high end"
individuals, general population, high exposure group(s), children, susceptible
populations.
- How was the central tendency estimate developed? What factors and/or methods
were used in developing this estimate?
- How was the high-end estimate developed?
- Is there information on highly exposed subgroups? Who are they? What are their
levels of exposure? How are they accounted for in the assessment?
E. Is there reason to be concerned about cumulative or multiple exposures because of
ethnic, racial, or socioeconomic reasons?
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F. If adverse health effects have been observed in wildlife species, characterize wildlife
exposure by discussing the relevant issues as in A through E above.
G. Summarize exposure conclusions and discuss the following:
- results of different approaches, i.e., modeling, monitoring, probability
distributions;
- limitations of each, and the range of most reasonable values; and
- confidence in the results obtained, and the limitations to the results.
PART TWO
RISK CONCLUSIONS AND COMPARISONS
IV. Risk Conclusions
A. What is the overall picture of risk, based on the hazard identification, dose-response
and exposure characterizations?
B. What are the major conclusions and strengths of the assessment in each of the three
main analyses (i.e., hazard identification, dose-response, and exposure assessment)?
C. What are the major limitations and uncertainties in the three main analyses?
D. What are the science policy options in each of the three major analyses?
- What are the alternative approaches evaluated?
- What are the reasons for the choices made?
V. Risk Context
A. What are the qualitative characteristics of the hazard (e.g., voluntary vs. involuntary,
technological vs. natural, etc.)? Comment on findings, if any, from studies of risk
perception that relate to this hazard or similar hazards.
B. What are the alternatives to this hazard? How do the risks compare?
C. How does this risk compare to other risks?
1. How does this risk compare to other risks in this regulatory program, or other
similar risks that the EPA has made decisions about?
2. Where appropriate, can this risk be compared with past Agency decisions,
decisions by other federal or state agencies, or common risks with which people
may be familiar?
3. Describe the limitations of making these comparisons.
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D. Comment on significant community concerns which influence public perception of
risk.
VI. Existing Risk Information
Comment on other risk assessments that have been done on this chemical by EPA, other
federal agencies, or other organizations. Are there significantly different conclusions that
merit discussion?
VII. Other Information
Is there other information that would be useful to the risk manager or the public in this
situation that has not been described above?
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GUIDANCE
FOR
RISK CHARACTERIZATION
U.S. Environmental Protection Agency
Science Policy Council
February 1995
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CONTENTS
I. The Risk Assessment-Risk Management Interface
II. Risk Assessment and Risk Characterization
IE. Exposure and Risk Descriptors
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PREFACE
This guidance contains principles for developing and describing EPA risk assessments,
with a particular emphasis on risk characterization. The current document is an update of the
guidance issued with the Agency's 1992 policy (Guidance on Risk Characterization for Risk
Managers and Risk Assessors, February 26, 1992). The guidance has not been substantially
revised, but includes some clarifications and changes to give more prominence to certain issues,
such as the need to explain the use of default assumptions.
As in the 1992 policy, some aspects of this guidance focus on cancer risk assessment, but
the guidance applies generally to human health effects (e.g., neurotoxicity, developmental
toxicity) and, with appropriate modifications, should be used in all health risk assessments. This
document has not been revised to specifically address ecological risk assessment; however,
initial guidance for ecological risk characterization is included in EPA's Framework for
Ecological Risk Assessments (EPA/630/R-92/001). Neither does this guidance address in detail
the use of risk assessment information (e.g., information from the Integrated Risk Information
System (IRIS)) to generate site- or media-specific risk assessments. Additional program-
specific guidance will be developed to enable implementation of EPA's Risk Characterization
Policy. Development of such guidance will be overseen by the Science Policy Council and will
involve risk assessors and risk managers from across the Agency.
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I. THE RISK ASSESSMENT-RISK MANAGEMENT INTERFACE
Recognizing that for many people the term risk assessment has wide meaning, the National
Research Council's 1983 report on risk assessment in the federal government distinguished
between risk assessment and risk management.
"Broader uses of the term [risk assessment] than ours also embrace analysis of
perceived risks, comparisons of risks associated with different regulatory strategies,
and occasionally analysis of the economic and social implications of regulatory
decisions—functions that we assign to risk management (emphasis added). (1)
In 1984, EPA endorsed these distinctions between risk assessment and risk management for
Agency use (2), and later relied on them in developing risk assessment guidelines (3). In 1994,
the NRC reviewed the Agency's approach to and use of risk assessment and issued an extensive
report on their findings (4). This distinction suggests that EPA participants in the process can be
grouped into two main categories, each with somewhat different responsibilities, based on their
roles with respect to risk assessment and risk management.
A. Roles of Risk Assessors anal Risk Managers
Within the Risk Assessment category there is a group that develops chemical-specific risk
assessments by collecting, analyzing, and synthesizing scientific data to produce the hazard
identification, dose-response, and exposure assessment portion of the risk assessment and to
characterize risk. This group relies in part on Agency risk assessment guidelines to address
science policy issues and scientific uncertainties. Generally, this group includes scientists and
statisticians in the Office of Research and Development; the Office of Prevention, Pesticides
and Toxics and other program offices; the Carcinogen Risk Assessment Verification Endeavor
(CRAVE); and the Reference Dose (RfD) and Reference Concentration (RfC) Workgroups
Another group generates site- or media-specific risk assessments for use in regulation
development or site-specific decision-making. These assessors rely on existing databases (e.g.,
IRIS, ORD Health Assessment Documents, CRAVE and RfD/RfC Workgroup documents, and
program-specific toxicity information) and media- or site-specific exposure information in
developing risk assessments. This group also relies in part on Agency risk assessment
guidelines and program-specific guidance to address science policy issues and scientific
uncertainties. Generally, this group includes scientists and analysts in program offices, regional
offices, and the Office of Research and Development.
Risk managers, as a separate category, integrate the risk characterization with other
considerations specified in applicable statutes to make and justify regulatory decisions.
Generally, this group includes Agency managers and decision-makers. Risk managers also play
a role in determining the scope of risk assessments. The risk assessment process involves
regular interaction between risk assessors and risk managers, with overlapping responsibilities
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at various stages in the overall process. Shared responsibilities include initial decisions
regarding the planning and conduct of an assessment, discussions as the assessment develops,
decisions regarding new data needed to complete an assessment and to address significant
uncertainties. At critical junctures in the assessment, such consultations shape the nature of, and
schedule for, the assessment. External experts and members of the public may also play a role
in determining the scope of the assessment; for example, the public is often concerned about
certain chemicals or exposure pathways in the development of site-specific risk assessments.
B. Guiding Principles
The following guidance outlines principles for those who generate, review, use, and integrate
risk assessments for decision-making.
1. Risk assessors and risk managers should be sensitive to distinctions between risk
assessment and risk management.
The major participants in the risk assessment process have many shared responsibilities. Where
responsibilities differ, it is important that participants confine themselves to tasks in their areas
of responsibility and not inadvertently obscure differences between risk assessment and risk
management.
For the generators of the assessment distinguishing between risk assessment and risk
management means that scientific information is selected, evaluated, and presented without
considering issues such as cost, feasibility, or how the scientific analysis might influence the
regulatory or site-specific decision. Assessors are charged with (1) generating a credible,
objective, realistic, and scientifically balanced analyst; (2) presenting information on hazard,
dose-response, exposure and risk; and (3) explaining confidence in each assessment by clearly
delineating strengths, uncertainties and assumptions, along with the impacts of these factors
(e.g., confidence limits, use of conservative/non-conservative assumptions) on the overall
assessment. They do not make decisions on the acceptability of any risk level for protecting
public health or selecting procedures for reducing risks.
For users of the assessment and for decision-makers who integrate these assessments into
regulatory or site-specific decisions, the distinction between risk assessment and risk
management means refraining from influencing the risk description through consideration of
other factors—e.g., the regulatory outcome—and from attempting to shape the risk assessment
to avoid statutory constraints, meet regulatory objectives, or serve political purposes. Such
management considerations are often legitimate considerations for the overall regulatory
decision (see next principle), but they have no role in estimating or describing risk. However,
decision-makers and risk assessors participate in an Agency process that establishes policy
directions that determine the overall nature and tone of Agency risk assessments and, as
appropriate, provide policy guidance on difficult and controversial risk assessment issues.
Matters such as risk assessment priorities, degree of conservatism, and acceptability of
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particular risk levels are reserved for decision-makers who are charged with making decisions
regarding protection of public health.
2. The risk assessment product, that is, the risk characterization, is only one of several
kinds of information used for regulatory decision-making.
Risk characterization, the last step in risk assessment, is the starting point for risk management
considerations and the foundation for regulatory decision-making, but it is only one of several
important components in such decisions. As the last step in risk assessment, the risk
characterization identifies and highlights the noteworthy risk conclusions and related
uncertainties. Each of the environmental laws administered by EPA calls for consideration of
other factors at various stages in the regulatory process. As authorized by different statutes,
decision-makers evaluate technical feasibility (e.g., treatability, detection limits), economic,
social, political, and legal factors as part of the analysis of whether or not to regulate and, if so,
to what extent. Thus, regulatory decisions are usually based on a combination of the technical
analysis used to develop the risk assessment and information from other fields.
For this reason, risk assessors and managers should understand that the regulatory decision is
usually not determined solely by the outcome of the risk assessment. For example, a regulatory
decision on the use of a particular pesticide considers not only the risk level to affected
populations, but also the agricultural benefits of its use that may be important for the nation's
food supply. Similarly, assessment efforts may produce an RfD for a particular chemical, but
other considerations may result in a regulatory level that is more or less protective than the RfD
itself.
For decision-makers, this means that societal considerations (e.g., costs and benefits) that, along
with the risk assessment, shape the regulatory decision should be described as fully as the
scientific information set forth in the risk characterization. Information on data sources and
analyses, their strengths and limitations, confidence in the assessment, uncertainties, and
alternative analyses are as important here as they are for the scientific components of the
regulatory decision. Decision-makers should be able to expect, for example, the same level of
rigor from the economic analysis as they receive from the risk analysis. Risk management
decisions involve numerous assumptions and uncertainties regarding technology, economics
and social factors, which need to be explicitly identified for the decision-makers and the public.
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II. RISK CHARACTERIZATION
A. Defining Risk Characterization in the Context of Risk Assessment
EPA risk assessment principles and practices draw on many sources. Obvious sources include
the environmental laws administered by EPA, the National Research Council's 1983 report on
risk assessment (1), the Agency's Risk Assessment Guidelines (3), and various program specific
guidance (e.g., the Risk Assessment Guidance for Superfund). Twenty years of EPA experience
in developing, defending, and enforcing risk assessment-based regulation is another. Together
these various sources stress the importance of a clear explanation of Agency processes for
evaluating hazard, dose-response, exposure, and other data that provide the scientific
foundation for characterizing risk.
This section focuses on two requirements for full characterization of risk. First, the
characterization should address qualitative and quantitative features of the assessment. Second,
it should identify the important strengths and uncertainties in the assessment as part of a
discussion of the confidence in the assessment. This emphasis on a full description of all
elements of the assessment draws attention to the importance of the qualitative, as well as the
quantitative, dimensions of the assessment. The 1983 NRC report carefully distinguished
qualitative risk assessment from quantitative assessments, preferring risk statements that are not
strictly numerical.
The term risk assessment is often given narrower and broader meanings than we have adopted
here. For some observers, the term is synonymous with quantitative risk assessment and
emphasizes reliance on numerical results. Our broader definition includes quantification, but
also includes qualitative expressions of risk. Quantitative estimates of risk are not always
feasible, and they may be eschewed by agencies for policy reasons. (1)
EPA's Exposure Assessment Guidelines define risk characterization as the final step in the risk
assessment process that:
Integrates the individual characterizations from the hazard identification, dose-
response, and exposure assessments;
• Provides an evaluation of the overall quality of the assessment and the degree of
confidence the authors have in the estimates of risk and conclusions drawn;
Describes risks to individuals and populations in terms of extent and severity of
probable harm; and
• Communicates results of the risk assessment to the risk manager. (5)
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Particularly critical to full characterization of risk is a frank and open discussion of the
uncertainty in the overall assessment and in each of its components. The uncertainty discussion
is important for several reasons.
1. Information from different sources carries different kinds of uncertainty and knowledge of
these differences is important when uncertainties are combined for characterizing risk.
2. The risk assessment process, with management input, involves decisions regarding the
collection of additional data (versus living with uncertainly); in the risk characterization, a
discussion of the uncertainties will help to identify where additional information could
contribute significantly to reducing uncertainties in risk assessment.
3. A clear and explicit statement of the strengths and limitations of a risk assessment
requires a clear and explicit statement of related uncertainties.
A discussion of uncertainty requires comment on such issues as the quality and quantity of
available data, gaps in the data base for specific chemicals, quality of the measured data, use of
default assumptions, incomplete understanding of general biological phenomena, and scientific
judgments or science policy positions that were employed to bridge information gaps.
In short, broad agreement exists on the importance of a full picture of risk, particularly
including a statement of confidence in the assessment and the associated uncertainties. This
section discusses information content and uncertainty aspects of risk characterization, while
Section HI discusses various descriptors used in risk characterization.
B. Guiding Principles
1. The risk characterization integrates the information from the hazard identification,
dose-response, and exposure assessments, using a combination of qualitative
information, quantitative information, and information regarding uncertainties.
Risk assessment is based on a series of questions that the assessor asks about the data and the
implications of the data for human risk. Each question calls for analysis and interpretation of
the available studies, selection of the data that are most scientifically reliable and most relevant
to the problem at hand, and scientific conclusions regarding the question presented. As
suggested below, because the questions and analyses are complex, a complete characterization
includes several different kinds of information, carefully selected for reliability and relevance.
a. Hazard Identification—What is known about the capacity of an environmental agent
for causing cancer (or other adverse effects) in humans and laboratory animals?
Hazard identification is a qualitative description based on factors such as the kind and quality of
data on humans or laboratory animals, the availability of ancillary information (e.g., structure-
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activity analysis, genetic toxicity, pharmacokinetics) from other studies, and the weight-of-the-
evidence from all of these data sources. For example, to develop this description, the issues
addressed include:
1) the nature, reliability, and consistency of the particular studies in humans and in
laboratory animals;
2) the available information on the mechanistic basis for activity; and
3) experimental animal responses and their relevance to human outcomes.
These issues make clear that the task of hazard identification is characterized by describing the
full range of available information and the implications of that information for human health.
b. Dose-Response Assessment—What is known about the biological mechanisms and
dose-response relationships underlying any effects observed in the laboratory or
epidemiology studies providing data for the assessment?
The dose-response assessment examines quantitative relationships between exposure (or dose)
and effects in the studies used to identify and define effects of concern. This information is later
used along with "real world" exposure information (see below) to develop estimates of the
likelihood of adverse effects in populations potentially at risk. It should be noted that, in
practice, hazard identification for developmental toxicity and other non-cancer health effects is
usually done in conjunction with an evaluation of dose-response relationships, since the
determination of whether there is a hazard is often dependent on whether a dose response
relationship is present. (6) Also, the framework developed by EPA for ecological risk
assessment does not distinguish between hazard identification and dose-response assessment,
but rather calls for a "characterization of ecological effects." (7)
Methods for establishing dose-response relationships often depend on various assumptions used
in lieu of a complete database, and the method chosen can strongly influence the overall
assessment. The Agency's risk assessment guidelines often identify so-called "default
assumptions" for use in the absence of other information. The risk assessment should pay
careful attention to the choice of a high-to-low dose extrapolation procedure. As a result, an
assessor who is characterizing a dose-response relationship considers several key issues:
1) the relationship between extrapolation models selected and available information on
biological mechanisms;
2) how appropriate data sets were selected from those that show the range of possible
potencies both in laboratory animals and humans;
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3) the basis for selecting interspecies dose scaling factors to account for scaling doses
from experimental animals to humans;
4) the correspondence between the expected route(s) of exposure and the exposure
route(s) utilized in the studies forming the basis of the dose-response assessment, as
well as the interrelationships of potential effects from different exposure routes;
5) the correspondence between the expected duration of exposure and the exposure
durations in the studies used in forming the basis of the dose-response assessment,
e.g., chronic studies would be used to assess long-term, cumulative exposure
concentrations, while acute studies would be used in assessing peak levels of
exposure; and
6) the potential for differing susceptibilities among population subgroups.
The Agency's Integrated Risk Information System (IRIS) is a repository for such information
for EPA. EPA program offices also maintain program-specific databases, such as the OSWER.
Health Effects Assessment Summary Tables (HEAST). IRIS includes data summaries
representing Agency consensus on specific chemicals, based on a careful review of the
scientific issues listed above. For specific risk assessments based on data from any source, risk
assessors should carefully review the information presented, emphasizing confidence in the data
and uncertainties (see subsection 2 below). Specifically, when IRIS data are used, the IRIS
statement of confidence should be included as an explicit part of the risk characterization for
hazard and dose-response information.
c. Exposure Assessment—What is known about the principal paths, patterns, and
magnitudes of human exposure and numbers of persons who may be exposed?
The exposure assessment examines a wide range of exposure parameters pertaining to the
environmental scenarios of people who may be exposed to the agent under study. The
information considered for the exposure assessment includes monitoring studies of chemical
concentrations in environmental media, food, and other materials; modeling of environmental
fate and transport of contaminants; and information on different activity patterns of different
population subgroups. An assessor who characterizes exposure should address several issues:
1) The basis for the values and input parameters used for each exposure scenario. If the
values are based on data, there should be a discussion of the quality, purpose, and
representativeness of the database. For monitoring data, there should be a discussion
of the data quality objectives as they are relevant to risk assessment, including the
appropriateness of the analytical detection limits. If models are applied, the
appropriateness of the models and information on their validation should be
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presented. When assumptions are made, the source and general logic used to develop
the assumptions (e.g., program guidance, analogy, professional judgment) should be
described.
2) The confidence in the assumptions made about human behavior and the relative
likelihood of the different exposure scenarios.
3) The major factor or factors (e.g., concentration, body uptake, duration/frequency of
exposure) thought to account for the greatest uncertainty in the exposure estimate,
due either to sensitivity or lack of data.
4) The link between the exposure information and the risk descriptors discussed in
Section HI of this Appendix. Specifically, the risk assessor needs to discuss the
connection between the conservatism or non-conservatism of the data/assumptions
used in the scenarios and the choice of descriptors.
5) Other information that may be important for the particular risk assessment. For
example, for many assessments, other sources and background levels in the
environment may contribute significantly to population exposures and should be
discussed.
2. The risk characterization includes a discussion of uncertainty and variability.
In the risk characterization, conclusions about hazard and dose response are integrated with
those from the exposure assessment. In addition, confidence about these conclusions, including
information about the uncertainties associated with each aspect of the assessment in the final
risk summary, is highlighted. In the previous assessment steps and in the risk characterization,
the risk assessor must distinguish between variability and uncertainty.
Variability arises from true heterogeneity in characteristics such as dose-response differences
within a population, or differences in contaminant levels in the environment. The values of
some variables used in an assessment change with time and space, or across the population
whose exposure is being estimated. Assessments should address the resulting variability in
doses received by members of the target population. Individual exposure, dose, and risk can
vary widely in a large population. The central tendency and high end individual risk descriptors
(discussed in Section HI below) are intended to capture the variability in exposure, lifestyles,
and other factors that lead to a distribution of risk across a population.
Uncertainty, on the other hand, represents lack of knowledge about factors such as adverse
effects or contaminant levels which may be reduced with additional study. Generally, risk
assessments carry several categories of uncertainty, and each merits consideration.
Measurement uncertainty refers to the usual error that accompanies scientific measurements—
standard statistical techniques can often be used to express measurement uncertainty. A
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substantial amount of uncertainty is often inherent in environmental sampling, and assessments
should address these uncertainties. There are likewise uncertainties associated with the use of
scientific models, e.g., dose-response models, models of environmental fate and transport.
Evaluation of model uncertainty would consider the scientific basis for the model and available
empirical validation.
A different kind of uncertainty stems from data gaps—that is, estimates or assumptions used in
the assessment. Often, the data gap is broad, such as the absence of information on the effects
of exposure to a chemical on humans or on the biological mechanism of action of an agent. The
risk assessor should include a statement of confidence that reflects the degree to which the risk
assessor believes that the estimates or assumptions adequately fill the data gap. For some
common and important data gaps, Agency or program-specific risk assessment guidance
provides default assumptions or values. Risk assessors should carefully consider all available
data before deciding to rely on default assumptions. If defaults are used, the risk assessment
should reference the Agency guidance that explains the default assumptions or values.
Often risk assessors and managers simplify discussion of risk issues by speaking only of the
numerical components of an assessment. That is, they refer to the alphanumeric weight-of-the-
evidence classification, unit risk, the risk-specific dose or the qx* for cancer risk, and the
RfD/RfC for health effects other than cancer, to the exclusion of other information bearing on
the risk case. However, since every assessment carries uncertainties, a simplified numerical
presentation of risk is always incomplete and often misleading. For this reason, the NRC (1)
and EPA risk assessment guidelines (2) call for "characterizing" risk to include qualitative
information, a related numerical risk estimate and a discussion of uncertainties, limitations, and
assumptions—default and otherwise.
Qualitative information on methodology, alternative interpretations, and working assumptions
(including defaults) is an important component of risk characterization. For example, specifying
that animal studies rather than human studies were used in an assessment tells others that the
risk estimate is based on assumptions about human response to a particular chemical rather than
human data. Information that human exposure estimates are based on the subjects' presence in
the vicinity of a chemical accident rather than tissue measurements defines known and
unknown aspects of the exposure component of the study.
Qualitative descriptions of this kind provide crucial information that augments understanding of
numerical risk estimates. Uncertainties such as these are expected in scientific studies and in
any risk assessment based on these studies. Such uncertainties do not reduce the validity of the
assessment. Rather, they should be highlighted along with other important risk assessment
conclusions to inform others fully on the results of the assessment.
In many cases, assessors must choose among available data, models, or assumptions in
estimating risks. Examining the impact of selected, plausible alternatives on the conclusions of
the assessment is an important part of the uncertainty discussion. The key words are "selected"
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and "plausible"; listing all alternatives to a particular assumption, regardless of their merits
would be superfluous. Generators of the assessment, using best professional judgment, should
outline the strengths and weaknesses of the plausible alternative approaches.1
An adequate description of the process of alternatives selection involves several aspects.
a. A rationale for the choice.
b. Discussion of the effects of alternatives selected on the assessment.
c. Comparison with other plausible alternatives, where appropriate.
The degree to which variability and uncertainty are addressed depends largely on the scope of
the assessment and the resources available. For example, the Agency does not expect an
assessment to evaluate and assess every conceivable exposure scenario for every possible
pollutant, to examine all susceptible populations potentially at risk, or to characterize every
possible environmental scenario to estimate the cause and effect relationships between exposure
to pollutants and adverse health effects. Rather, the discussion of uncertainty and variability
should reflect the type and complexity of the risk assessment, with the level of effort for
analysis and discussion of uncertainty corresponding to the level of effort for the assessment.
3. 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.
The risk assessment process calls for identifying and highlighting significant risk conclusions
and related uncertainties partly to assure full communication among risk assessors and partly to
assure that decision-makers are fully informed. Issues are identified by acknowledging
noteworthy qualitative and quantitative factors that make a difference in the overall assessment
of hazard and risk, and hence in the ultimate regulatory decision. The key word is
"noteworthy." Information that significantly influences the analysis is explicitly noted—in all
future presentations of the risk assessment and in the related decision. Uncertainties and
assumptions that strongly influence confidence in the risk estimate also require special
attention.
Numerical estimates should not be separated from the descriptive information that is integral to
risk characterization. Documents and presentations supporting regulatory or site-specific
decisions should include both the numerical estimate and descriptive information; in short
reports, this information can be abbreviated. Fully visible information assures that important
features of the assessment are immediately available at each level of review for evaluating
whether risks are acceptable or unreasonable.
JIn cases where risk assessments within an Agency program routinely address similar sets of alternatives,
program guidance may be developed to streamline and simplify the discussion of these alternatives.
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III. EXPOSURE ASSESSMENT AND RISK DESCRIPTORS
A. Presentation of Risk Descriptors
The results of a risk assessment are usually communicated to the risk manager in the risk
characterization portion of the assessment. This communication is often accomplished through
risk descriptors which convey information and answer questions about risk, each descriptor
providing different information and insights. Exposure assessment plays a key role in
developing these risk descriptors since each descriptor is based in part on the exposure
distribution within the population of interest.
The following guidance outlines the different descriptors in a convenient order that should not
be construed as a hierarchy of importance. These descriptors should be used to describe risk in
a variety of ways for a given assessment, consistent with the assessment's purpose, the data
available, and the information the risk manager needs. Use of a range of descriptors instead of a
single descriptor enables Agency programs to present a picture of risk that corresponds to the
range of different exposure conditions encountered for most environmental chemicals. This
analysis, in turn, allows risk managers to identify populations at greater and lesser risk and to
shape regulatory solutions accordingly.
Agency risk assessments will be expected to address or provide descriptions of (1) individual
risk that include the central tendency and high end portions of the risk distribution, (2)
population risk, and (3) important subgroups of the population, such as highly exposed or
highly susceptible groups. Assessors may also use additional descriptors of risk as needed when
these add to the clarity of the presentation. With the exception of assessments where particular
descriptors clearly do not apply, some form of these three types of descriptors should be
routinely developed and presented for Agency risk assessments.2 In other cases, where a
descriptor would be relevant, but the program lacks the data or methods to develop it, the
program office should design and implement a plan, in coordination with other EPA offices, to
meet these assessment needs. While gaps continue to exist, risk assessors should make their
best efforts to address each risk descriptor, and at a minimum, should briefly discuss the lack of
data or methods. Finally, presenters of risk assessment information should be prepared to
routinely answer questions by risk managers concerning these descriptors.
It is essential that presenters not only communicate the results of the assessment by addressing
each of the descriptors where appropriate, but that they also communicate their confidence that
these results portray a reasonable picture of the actual or projected exposures. This task will
Program-specific guidance will need to address these situations. For example, for site-specific
assessments, the utility and appropriateness of population risk estimates will be determined based on the available
data and program guidance.
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usually be accomplished by frankly commenting on the key assumptions and parameters that
have the greatest impact on the results, the basis or rationale for choosing these assumptions/
parameters, and the consequences of choosing other assumptions.
B. Relationship Between Exposure Descriptors and Risk Descriptors
In the risk assessment process, risk is estimated as a function of exposure, with the risk of
adverse effects increasing as exposure increases. Information on the levels of exposure
experienced by different members of the population is key to understanding the range of risks
that may occur. Risk assessors and risk managers should keep in mind, however, that exposure
is not synonymous with risk. Differences among individuals, in absorption rates, susceptibility,
or other factors mean that individuals with the same level of exposure may be at different levels
of risk. In most cases, the state of the science is not yet adequate to define distributions of
factors such as population susceptibility. The guidance principles below discuss a variety of risk
descriptors that primarily reflect differences in estimated exposure. If a full description of the
range of susceptibility in the population cannot be presented, an effort should be made to
identify subgroups that, for various reasons, may be particularly susceptible.
C. Guiding Principles
1. Information about the distribution of individual exposures is important to
communicating the results of a risk assessment.
The risk manager is generally interested in answers to questions such as the following:
Who are the people at the highest risk?
What risk levels are they subjected to?
What are they doing, where do they live, etc., that might be putting them at this
higher risk?
• What is the average risk for individuals in the population of interest?
Individual exposure and risk descriptors are intended to provide answers to these questions so
as to illuminate the risk management decisions that need to be made. In order to describe the
range of risks, both high end and central tendency descriptors are used to convey the variability
in risk levels experienced by different individuals in the population.
a. High end descriptor
For the Agency's purposes, high end risk descriptors are plausible estimates of the individual
risk for those persons at the upper end of the risk distribution. Given limitations in current
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understanding of variability in individuals' sensitivity to toxins, high end descriptors will
usually address high end exposure or dose (herein referred to as exposure for brevity). The
intent of these descriptors is to convey estimates of exposure in the upper range of the
distribution, but to avoid estimates which are beyond the true distribution. Conceptually, high
end exposure means exposure above about the 90th percentile of the population distribution,
but not higher than the individual in the population who has the highest exposure. When large
populations are assessed, a large number of individuals may be included within the "high end"
(e.g., above 90th or 95th percentile) and information on the range of exposures received by
these individuals should be presented.
High end descriptors are intended to estimate the exposures that are expected to occur in small,
but definable, "high end" segments of the subject population.3 The individuals with these
exposures may be members of a special population segment or individuals in the general
population who are highly exposed because of the inherent stochastic nature of the factors
which give rise to exposure. Where differences in sensitivity can be identified within the
population, high end estimates addressing sensitive individuals or subgroups can be developed.
In those few cases in which the complete data on the population distributions of exposures and
doses are available, high end exposure or dose estimates can be represented by reporting
exposures or doses at a set of selected percentiles of the distributions, such as the 90th, 95th,
and 98th percentile. High end exposures or doses, as appropriate, can then be used to calculate
high end risk estimates.
In the majority of cases where the complete distributions are not available, several methods
help estimate a high end exposure or dose. If sufficient information about the variability in
chemical concentrations, activity patterns, or other factors are available, the distribution may be
estimated through the use of appropriate modeling (e.g., Monte Carlo simulation or parametric
statistical methods). The determination of whether available information is sufficient to support
the use of probabilistic estimation methods requires careful review and documentation by the
risk assessor. If the input distributions are based on limited data, the resulting distribution
should be evaluated carefully to determine whether it is an improvement over more traditional
estimation techniques. If a distribution is developed, it should be described with a series of
percentiles or population frequency estimates, particularly in the high end range. The assessor
and risk manager should be aware, however, that unless a great deal is known about exposures
and doses at the high end of the distribution, these estimates will involve considerable
3High end estimates focus on estimates of exposure in the exposed populations. Bounding estimates, on the
other hand, are constructed to be equal to or greater than the highest actual risk in the population (or the highest risk
that could be expected in a future scenario). A "worst case scenario" refers to a combination of events and
conditions such that, taken together, produces the highest conceivable risk. Although it is possible that such an
exposure, dose, or sensitivity combination might occur in a given population of interest, the probability of an
individual receiving this combination of events and conditions is usually small, and often so small that such a
combination will not occur in a particular, actual population.
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uncertainty which the exposure assessor will need to describe. Note that in this context, the
probabilistic analysis addresses variability of exposure in the population. Probabilistic
techniques may also be applied to evaluate uncertainty in estimates (see section 5, below).
However, it is generally inappropriate to combine distributions reflecting both uncertainty and
variability to get a single overall distribution. Such a result is not readily interpretable for the
concerns of environmental decision-making.
If only limited information on the distribution of the exposure or dose factors is available, the
assessor should approach estimating the high end by identifying the most sensitive variables
and using high end values for a subset of these variables, leaving others at their central values.4
In doing this, the assessor needs to avoid combinations of parameter values that are inconsistent
(e.g., low body weight used in combination with high dietary intake rates), and must keep in
mind the ultimate objective of being within the distribution of actual expected exposures and
doses, and not beyond it.
If very little data are available on the ranges for the various variables, it will be difficult to
estimate exposures or doses and associated risks in the high end with much confidence. One
method that has been used in such cases is to start with a bounding estimate and "back off the
limits used until the combination of parameter values is, in the judgment of the assessor, within
the distribution of expected exposure, and still lies within the upper 10% of persons exposed.
Obviously, this method results in a large uncertainty and requires explanation.
b. Central tendency descriptor
Central tendency descriptors generally reflect central estimates of exposure or dose. The
descriptor addressing central tendency may be based on either the arithmetic mean exposure
(average estimate) or the median exposure (median estimate), either of which should be clearly
labeled. The average estimate, used to approximate the arithmetic mean, can often be derived
by using average values for all the exposure factors.5 It does not necessarily represent a
particular individual on the distribution. Because of the skewness of typical exposure profiles,
the arithmetic mean may differ substantially from the median estimate (i.e., 50th percentile
estimate, which is equal to the geometric mean for a log normal distribution). The selection of
which descriptor(s) to present in the risk characterization will depend on the available data and
the goals of the assessment. When data are limited, it may not be possible to construct true
4Maximizing all variables will in virtually all cases result in an estimate that is above the actual values seen
in the population. When the principal parameters of the dose equation, e.g., concentration (appropriately integrated
over time), intake rate, and duration, are broken out into sub-components, it may be necessary to use maximum
values for more than two of these sub-component parameters depending on a sensitivity analysis.
5This holds true when variables are added (e.g., exposures by different routes) or when independent
variables are multiplied (e.g., concentration x intake). However, it would be incorrect for products of correlated
variables, variables used as divisors, or for formulas involving exponents.
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median or mean estimates, but it is still possible to construct estimates of central tendency. The
discussion of the use of probabilistic techniques in Section l(a) above also applies to estimates
of central tendency.
2. Information about population exposure leads to another important way to describe
risk.
Population risk refers to an assessment of the extent of harm for the population as a whole. In
theory, it can be calculated by summing the individual risks for all individuals within the
subject population. This task, of course, requires a great deal more information than is
normally, if ever, available.
The kinds of questions addressed by descriptors of population risk include the following:
• How many cases of a particular health effect might be probabilistically estimated in
this population for a specific time period?
For non-carcinogens, what portion of the population is within a specified range of
some reference level; e.g., exceedance of the RfD (a dose), the RfC (a concentration),
or other health concern level?
For carcinogens, what portion of the population is above a certain risk level, such as
1Q-6?
These questions can lead to two different descriptors of population risk.
a. Probabilistic number of cases
The first descriptor is the probabilistic number of health effect cases estimated in the population
of interest over a specified time period. This descriptor can be obtained either by (a) summing
the individual risks over all the individuals in the population, e.g. using an estimated
distribution of risk in the population, when such information is available, or (b) through the use
of a risk model that assumes a linear non-threshold response to exposure, such as many
carcinogenic models. In these calculations, data will typically be available to address variability
in individual exposures. If risk varies linearly with exposure, multiplying the mean risk by the
population size produces an estimate of the number of cases.6 At the present time, most cancer
potency values represent plausible upper bounds on risk. When such a value is used to estimate
6However, certain important cautions apply (see EPA's Exposure Assessment Guidelines). Also, this is not
appropriate for non-carcinogenic effects or for other types of cancer models. For non-linear cancer models, an
estimate of population risk must be calculated using the distribution of individual risks
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numbers of cancer cases, it is important to understand that the result is also an upper bound. As
with other risk descriptors, this approach may not adequately address sensitive subgroups for
which different dose-response curve or exposure estimates might be needed.
Obviously, the more information one has, the more certain the estimate of this risk descriptor,
but inherent uncertainties in risk assessment methodology place limitations on the accuracy of
the estimate. The discussion of uncertainty involved in estimating the number of cases should
indicate that this descriptor is not to be confused with an actuarial prediction of cases in the
population (which is a statistical prediction based on a great deal of empirical data).
In general, it should be recognized that when small populations are exposed, population risk
estimates may be very small. For example, if 100 people are exposed to an individual lifetime
cancer risk of 10"4, the expected number of cases is 0.01. In such situations, individual risk
estimates will usually be a more meaningful parameter for decision-makers.
b. Estimated percentage of population with risk greater than some level
For non-cancer effects, we generally have not developed the risk assessment techniques to the
point of knowing how to add risk probabilities, so a second descriptor is usually more
appropriate: An estimate of the percentage of the population, or the number of persons, above a
specified level of risk or within a specified range of some reference level, e.g., exceedance of
the RfD or the RfC, LOAEL or other specific level of interest. This descriptor must be obtained
through measuring or simulating the population distribution.
3. Information about the distribution of exposure and risk for different subgroups of
the population are important components of a risk assessment.
A risk manager might also ask questions about the distributor of the risk burden among various
segments of the subject population such as the following: How do exposure and risk impact
various subgroups?; and, what is the population risk of a particular subgroup? Questions about
the distribution of exposure and risk among such population segments require additional risk
descriptors.
a. Highly exposed
Highly exposed subgroups can be identified, and where possible, characterized and the
magnitude of risk quantified. This descriptor is useful when there is (or is expected to be) a
subgroup experiencing significantly different exposures or doses from that of the larger
population. These sub-populations may be identified by age, sex, lifestyle, economic factors, or
other demographic variables. For example, toddlers who play in contaminated soil and high fish
consumers represent sub-populations that may have greater exposures to certain agents.
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b. Highly susceptible
Highly susceptible subgroups can also be identified, and if possible, characterized and the
magnitude of risk quantified. This descriptor is useful when the sensitivity or susceptibility to
the effect for specific subgroups is (or is expected to be) significantly different from that of the
larger population. In order to calculate risk for these subgroups, it will sometimes be necessary
to use a different dose-response relationship; e.g., upon exposure to a chemical, pregnant
women, elderly people, children, and people with certain illnesses may each be more sensitive
than the population as a whole. For example, children are thought to be both highly exposed
and highly susceptible to the effects of environmental lead. A model has been developed that
uses data on lead concentrations in different environmental media to predict the resulting blood
lead levels in children. Federal agencies are working together to develop specific guidance on
blood lead levels that present risks to children
It is important to note, however, that the Agency's current methodologies for developing
reference doses and reference concentrations (RfDs and RfCs) are designed to protect sensitive
populations. If data on sensitive human populations are available (and there is confidence in the
quality of the data), then the RfD is set at the dose level at which no adverse effects are
observed in the sensitive population (e.g., RfDs for fluoride and nitrate). If no such data are
available (for example, if the R is developed using data from humans of average or unknown
sensitivity), then an additional 10-fold factor is used to account for variability between the
average human response and the response of more sensitive individuals.
Generally selection of the population segments is a matter of either a priori interest in the
subgroup (e.g., environmental justice considerations), in which case the risk assessor and risk
manager can jointly agree on which subgroups to highlight, or a matter of discovery of a
sensitive or highly exposed subgroup during the assessment process. In either case, once
identified, the subgroup can be treated as a population in itself, and characterized in the same
way as the larger population using the descriptors for population and individual risk.
4. Situation-specific information adds perspective on possible future events or
regulatory options.
"What if...?" questions can be used to examine candidate risk management options. For
example, consider the following:
• What if a pesticide applicator applies this pesticide without using protective
equipment?
• What if this site becomes residential in the future?
What risk level will occur if we set the standard at 100 ppb?
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Answering these "What if...?" questions involves a calculation of risk based on specific
combinations of factors postulated within the assessment.7 The answers to these "What if...?"
questions do not, by themselves, give information about how likely the combination of values
might be in the actual population or about how many (if any) persons might be subjected to the
potential future risk. However, information on the likelihood of the postulated scenario would
also be desirable to include in the assessment.
When addressing projected changes for a population (either expected future developments or
consideration of different regulatory options), it is usually appropriate to calculate and consider
all the risk descriptors discussed above. When central tendency or high end estimates are
developed for a future scenario, these descriptors should reflect reasonable expectations about
future activities. For example, in site-specific risk assessments, future scenarios should be
evaluated when they are supported by realistic forecasts of future land use, and the risk
descriptors should be developed within that context.
5. An evaluation of the uncertainty in the risk descriptors is an important component
of the uncertainty discussion in the assessment.
Risk descriptors are intended to address variability of risk within the population and the overall
adverse impact on the population. In particular, differences between high end and central
tendency estimates reflect variability in the population, but not the scientific uncertainty
inherent in the risk estimates. As discussed above, there will be uncertainty in all estimates of
risk. These uncertainties can include measurement uncertainties, modeling uncertainties, and
assumptions to fill data gaps. Risk assessors should address the impact of each of these factors
on the confidence in the estimated risk values.
Both qualitative and quantitative evaluations of uncertainty provide useful information to users
of the assessment. The techniques of quantitative uncertainty analysis are evolving rapidly and
both the SAB (8) and the NRC (4) have urged the Agency to incorporate these techniques into
its risk analyses. However, it should be noted that a probabilistic assessment that uses only the
assessor's best estimates for distributions of population variables addresses variability, but not
uncertainty. Uncertainties in the estimated risk distribution need to be separately evaluated.
Some programs routinely develop future scenarios as part of developing a risk assessment. Program-
specific guidance may address future scenarios in more detail than they are described here.
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REFERENCES
1. National Research Council. Risk Assessment in the Federal Government: Management
the Process. 1983.
2. U.S. EPA. Risk Assessment and Management: Framework for Decision Making. 1984.
3. U.S. EPA. "Risk Assessment Guidelines." 51 Federal Register, 33992-34054, September
24, 1986.
4. National Research Council. Science and Judgement in Risk Assessment. 1994.
5. U.S. EPA. "Guidelines for Exposure Assessment." 57 Federal Register, 22888-22938,
May 29,1992.
6. U.S. EPA. "Guidelines for Developmental Toxicity Risk Assessment." 56 Federal
Register, 67398-63826, December 5, 1991.
7. U.S. EPA. Framework for Ecological Risk Assessment. 1992.
8. Loehr, R. A., and Matanoski, G.M., Letter to Carol M. Browner, EPA Administrator, Re:
Quantitative Uncertainty Analysis for Radiological Assessments. EPA Science Advisory
Board, July 23, 1993 (EPA-SAB-RAC-COM-93-006).
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APPENDIX E
ADDITIONAL DEVELOPMENTAL TOXICITY ISSUES
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APPENDIX E
APPENDIX E
ADDITIONAL DEVELOPMENTAL TOXICITY ISSUES
Several chemicals, including lead, PCBs, methylmercury, and some pharma-
ceuticals, are known to cause developmental toxicity in humans. This information
comes from large-scale poisoning incidents that resulted in serious developmental
effects in a large number of offspring. Human dose-response studies cannot be
carried out with planned dosing for developmental toxicants. However,
developmental toxicity studies have been carried out on many environmental
contaminants in animals. Many of these have yielded positive results (U.S. EPA,
1991). It is difficult to specifically interpret the dose-response relationship between
effects in animal studies and anticipated observable effects in the human
population. Research has been conducted to evaluate the relationship between
known human developmental toxicants and animal testing results; many
similarities in response were found. Alternatively, chemicals that caused develop-
mental effects in animals were studied for effects in humans. These evaluations
have yielded mixed results. It has been theorized that the lack of concurrence in
results may be due in part to the limited nature of the human data differences in
exposure route and the timing and duration of exposure (U.S. EPA, 1991). Further
analysis has indicated that:
The minimally effective dose for the most sensitive animal species
was generally higher than that for humans usually within 10-fold of
the human effective dose, but sometimes was 100 times or more
higher (U.S. EPA, 1991).
The Guidelines go on to state that:
Thus, the experimental animal data were generally predictive of
adverse developmental effects in humans, but in some cases, the
administered dose or exposure level required to achieve these
adverse effects was much higher than the effective dose in humans.
(U.S. EPA, 1991)
A number of assumptions are made in approaching developmental toxicity risk
assessment in the absence of specific information:
• Adverse effects in experimental animals may pose a hazard to humans.
• The four manifestations of developmental toxicity (death, structural abnor-
malities, growth alterations, and functional deficits) are all of concern rather
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APPENDIX E
than only malformations and death, which were the primary effects
considered in the past.
• The type of developmental effects seen in animals is not necessarily the
same as that produced in humans.
• The most appropriate species is used to estimate human risk when data are
available (e.g., pharmacokinetic). In the absence of such data, the most
sensitive species is used.
• A threshold is assumed based on the capacity of the developing organism
to repair or compensate for some amount of damage (U.S. EPA, 1991).
Although it is assumed there is a threshold for developmental toxicity, EPA has
stated that:
... a threshold for a population of individuals may or may not exist
because of other endogenous or exogenous factors that may
increase the sensitivity of some individuals in the population (U.S.
EPA, 1991).
The Agency is currently sponsoring research to better characterize the dose-
response relationship for developmental toxicants. This includes an evaluation of
the threshold concept (U.S. EPA, 1991). The process of risk assessment, as
recommended in the 1991 EPA guidelines, generally follows the four-step process
described in this document. However, hazard identification and dose-response
evaluation are combined in the developmental toxicity guidelines because "the
determination of hazard is often dependent on whether a dose-response
relationship is present" (U.S. EPA, 1991).
E. 1 DEFINITIONS
There is no one consistent definition of developmental toxicity (U.S. EPA, 1986a).
Developmental toxicity may include the range of effects from early pregnancy loss
to cognitive disorders detectable only long after birth. The severity of develop-
mental effects ranges from minor alterations in enzyme levels, with no known
associated pathology, to death. Developmental toxicity also encompasses health
endpoints having genetic and nongenetic bases. EPA's 1986 guidelines (U.S.
EPA, 1986a) provide useful definitions that are used in this document to classify
different types of developmental effects and to define the scope of effects
included under the overall heading of developmental effects.
• Developmental Toxicology—The study of adverse effects on the
developing organism that may result from exposure prior to conception
(either parent), during prenatal development, or postnatally to the time of
sexual maturation. Adverse developmental effects may be detected at any
point in the lifespan of the organism. The major manifestations of
developmental toxicity include: (1) death of the developing organism, (2 )
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APPENDIX E
structural abnormality, (3) altered growth (defined below), and (4) functional
deficiency.
• Functional Developmental Toxicology—The study of alterations or delays
in the physiological and/or biochemical functioning of the individual during
critical pre- or postnatal development periods.
• Embryotoxicity and Fetotoxicity—Any toxic effect on the conceptus as a
result of prenatal exposure. The distinguishing feature between the two
terms is the stage of development during which the injury occurs (the
embryonic stage lasts until approximately 8 weeks postconception followed
by the fetal stage). The terms include malformations and variations, altered
growth, and in utero death.
• Altered Growth—An alteration in offspring organ or body weight or size.
These alterations may or may not be accompanied by a change in crown-
rump length and/or in skeletal ossification. Altered growth can be induced
at any stage of development and may be reversible or may result in a
permanent change.
• Malformations—Permanent structural changes that may adversely affect
survival, development, or function. The term teratogenicity is used to
describe only structural abnormalities.
• Variations—Divergences beyond the usual range of structural constitution
that may not adversely affect survival or health. Distinguishing between
variations and malformations is difficult because responses form a
continuum from normal to extremely deviant. (U.S. EPA, 1986a, 1991).
Other terminology is often used (e.g., anomalies, deformations, and aberrations)
but definitions may vary.
For purposes of this guidance document, the definition of developmental
toxicology given above will be used to describe the range of effects considered
in this section. This provides a broad scope for evaluation of developmental
effects, including those resulting from both prenatal and preconception exposures
and effects that are observable pre- and postnatally. This section does not
include a discussion of reproductive system effects (i.e., damage to the
reproductive system), such as sterility, that result from exposure during adulthood
and that may prevent conception from occurring but that do not affect the
development of another individual. This type of toxicity is included under the
Chronic Toxicity heading in each profile in Section 5.
Carcinogenic effects occurring prior to adulthood may be considered
developmental effects under some circumstances. These can be evaluated using
the methods described in the previous section on carcinogenicity in keeping with
EPA recommendations (U.S. EPA, 1986b, 1996) and, similarly, mutagenic effects
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APPENDIX E
can be evaluated using criteria discussed in Guidelines for Mutagenicity Risk
Assessment (U.S. EPA, 1986c), as described in Appendix D.
E.2 SPECIAL ISSUES IN EVALUATING DEVELOPMENTAL TOXICANTS
Studies of developmental toxicants that are most useful in quantitative risk
assessment include human epidemiological studies and animal toxicology studies.
Epidemiological studies have been conducted on very few chemicals. Animal
studies, which are more readily available, pose problems related to interspecies
extrapolation (see statements in Sections 2.3.5 and 5 regarding uncertainty). The
Guidelines forthe Health Assessment of Suspect Developmental Toxicants (U.S.
EPA, 1991) provides guidance on evaluating various types of developmental
toxicity studies.
Some aspects of the evaluation of developmental toxicity studies differ from the
approaches and data that would be sought from most other types of toxicity
studies. One area of concern is the need to ascertain overall reproductive
performance, not only adverse effects on developing individuals. Exposure to a
toxicant often results in developmental damage at a very early stage of growth.
This may prevent implantation or lead to very early fetal loss. Such losses are
usually only detectable in animal studies by comparing the number of individuals
per litter or the number of litters produced to the same outcomes in control
populations. Very early losses are often absorbed and are not identifiable via
other means. In human studies such losses are not usually identified, although
prospective studies have used the monitoring of pregnancy markers, such as
human chorionic gonadotropic (HCG) hormone, to identify very early post-
implantation pregnancy losses (see U.S. EPA, 1991, for further discussion).
Another area of concern in developmental toxicity studies that is not usually of
significant interest in other types of toxicity studies is the importance of weight
changes. According to the federal guidelines, "A change in offspring body weight
is a sensitive indicator of developmental toxicity . . ." (U.S. EPA, 1991). A
relatively small weight change is not generally considered important in
toxicological studies of adult subjects; however, this is considered an important
effect during development. For example, the human corollary to decreased weight
in animals may be low birth weight, although this cannot be directly implied from
animal studies. Low birth weight in infants is a significant and often serious public
health problem. Weight gain or loss may also be organ-specific and may be
indicative of organ toxicity. For example, decreased brain weight may be
indicative of retarded or neurological development.
An issue that is often raised in developmental toxicity studies is maternal toxicity.
Although some researchers have suggested that the presence of maternal toxicity
undermines the validity of results observed in offspring, some level of maternal
toxicity should be observed in this type of study at the high end of the dose
regimen (U.S. EPA, 1991). The EPA health assessment guidelines describe
appropriate endpoints of maternal toxicity. One reason that identification of
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APPENDIX E
maternal toxicity is an important component of a developmental toxicity study is
that it can provide information on the likelihood of developing individuals being
more or less susceptible than adults to an agent. Agents that produce
developmental toxicity in offspring at doses that do not cause maternal toxicity are
of greatest concern because these dynamics suggest that developing individuals
are more sensitive or selectively affected (U.S. EPA, 1991). Those that produce
effects in parent and offspring at the same dose are also of concern; it should not
be assumed that offspring toxicity results from maternal toxicity because both may
be sensitive to the given dose level (U.S. EPA, 1991).
E.3 REFERENCES
U.S. EPA (U.S. Environmental Protection Agency). 1986a. Guidelines for the
health assessment of suspect developmental toxicants. Federal Register
51(185):34028-34040.
U.S. EPA (U.S. Environmental Protection Agency). 1986b. Guidelines for
carcinogen risk assessment. Federal Register 51(185):33992-34003.
U.S. EPA (U.S. Environmental Protection Agency). 1986c. Guidelines for
mutagenicity risk assessment. Federal Register 51(185):34006-34012.
U.S. EPA (U.S. Environmental Protection Agency). 1991. Guidelines for the
health assessment of suspect developmental toxicants. Federal Register
56:63798-63826.
U.S. EPA (U.S. Environmental Protection Agency). 1996. Proposed Guidelines
for Carcinogen Risk Assessment. EPA/600/P-92/003C, Office of Research
and Development, Washington, DC.
E-7
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APPENDIX F
SUMMARY OF LIMITS OF DETECTION FOR
THE RECOMMENDED TARGET ANALYTES
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APPENDIX F
Table F-1. Summary of Limits of Detection for the Recommended
Target Analytes3
Target Analyte Detection Limits'3 (ppb)
Metals
Arsenic (inorganic)0 5
Cadmiumd 5
Mercury8 1.3
Selenium' 17
Tributyltin9 2
Organochlorine Pesticides'
Chlordane (total) 1
c/s-Chlordane
frans-Chlordane
c/s-Nonachlor
frans-Nonachlor
Oxychlordane
DDT (Total)
4,4'-DDT 0.1
1,4'-DDT
4,4'-DDD
2,4'-DDD
4,4'-DDE
2,4'-DDE
Dicofol 1
Dieldrin 0.1
Endosulfan (Total) 5
Endosulfan I
Endosulfan II
Endrin 0.1
Heptachlor epoxide 0.1
Hexachlorobenzene 0.1
Lindane 0.1
Mirex 0.1
Toxaphene 3
Organophosphate Pesticides'
Chlorpyrifos 2
Diazinon 2
Disulfoton 2
Ethion 2
Turbufos 2
(continued)
F-3
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APPENDIX F
Table F-1 (continued)
Target Analyte Detection Limits'3 (ppb)
Chlorophenoxy Herbicides'
Oxyfluorfen 10
PAHsj 1 ppt
PCBs (Total Aroclors)h 20
Dioxins/Furans (Totalf 1 ppt
PAHs = Polycyclic aromatic hydrocarbons.
PCBs = Polychlorinated biphenyls.
a Detection limit provided for analysis of tissue on a wet weight basis.
b Limit of detection shown is lowest value identified. For further information, see Table
8-4, Volume 1, of this series.
c Analysis by hydride generation atomic absorption spectrophotometry (HAA) with
preconcentration (E. Crecelius, Battelle Pacific Northwest Laboratories, Marine
Sciences Laboratory, Sequim, WA, personal communication, July 1999).
d Analysis by graphite furnace atomic absorption spectrophotometry (GFAA).
e Analysis by cold vapor atomic absorption spectrophotometry (CVAA).
f Analysis by hydride generation on atomic absorption spectrophotometry (HAA).
9 Analysis by gas chromatography/flame photometric detection (GC/FPD) (E.
Crecelius, Battelle Pacific Northwest Laboratories, Marine Sciences Laboratory,
Sequim, WA, personal communication, July 1999).
h Analysis by gas chromatography/electron capture detection (GC/ECD), except where
otherwise noted. GC/ECD does not provide definitive compound identification, and
false positives due to interferences are commonly reported. Confirmation by an
alternative GC column phase (with ECD), or by GC/MS with selected ion monitoring,
is required for positive identification of PCBs, organochlorine pesticides, and
chlorophenoxy herbicides.
' Analysis by gas chromatography/nitrogen-phosphorus detection (GC/NPD).
' Analysis by gas chromatography/mass spectrometry (GC/MS). Detection limits of <1
ppb can be achieved using high-resolution gas chromatography/mass spectrometry
(HRGC/HRMS).
k Analysis by high-resolution GC/high-resolution mass spectrometry (HRGC/HRMS).
F-4
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