United States
Environmental Protection
Agency
Office of Water
(4305)
EPA 823-B-97-009
July 1997
&EPA Guidance for Assessing
Chemical Contaminant Data
for Use in Fish Advisories
Volume 2
Risk Assessment and Fish
Consumption Limits
Second Edition
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Guidance for Assessing Chemical Contaminant
Data for Use in Fish Advisories
Volume 2: Risk Assessment and Fish Consumption Limits
Second Edition
Office of Science and Technology
Office of Water
U.S. Environmental Protection Agency
Washington, DC
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UNITED STATES ENVIRONMENTAL PROTECTION AGENCY
WASHINGTON, D.C. 20460
Dear Colleagues:
OFFICE OF
WATER
The U.S. Environmental Protection Agency (EPA) is pleased to
announce the availability of the Second Edition of the document
titled Guidance for Assessing Chemical Contaminant Data For Use in
Fish Advisories Volume II: Risk Assessment and Fish Consumption
Limits(EPA 823-B-97-009). The purpose of the document is to •
provide the states, tribes and other interested parties with the
methods recommended by EPA for calculating consumption limits for
reducing health risks associated with contaminated fish and
shellfish. The Second Edition provides new information pertinent to
selected chemical contaminants as well as updated guidance on
methods for developing fish consumption advisories. Updated
information is provided on mercury, PCBs, PAHs, TBT and arsenic.
The document also contains a new section titled "Mapping Tools", an
expanded "Exposure Assessment" section, and revisions to sections
addressing the development and use of fish consumption limits.
The Second Edition is part of the USEPA's four volume series
of documents developed to provide guidance to professionals
responsible for assessing the health risks associated with exposure
to chemical contaminants in noncommercial fish and shellfish. This
manual is the Second Edition of Volume II in the four volume set of
documents prepared by the EPA. Volume I: Fish Sampling and
Analysis, First Edition was released in September 1993, with a
Second Edition released September 1995. Volume II: Risk Assessment
and Fish Consumption Limits was first published in June 1994.
Volume III: Risk Management was published in June 1996. Volume IV:
Risk Communication was published in March 1995.
These guidance documents are developed cooperatively with
State, Federal, Tribal and Local Government Agencies. All four of
these documents should be used together, as no single volume
addresses all of the topics necessary for developing fish
consumption advisories. Copies of the Second Edition of Volume II
(or the other Volumes) may be obtained by writing to the U.S.
Environmental Protection Agency, National Center For Environmental
Publications and Information, 11029 Kenwood Rd., Cincinnati, Ohio,
45242, or calling 513-489-8190. For questions related to the
development or use of this series of documents, please call Jeffrey
Bigler (202) 260-1305 of the EPA Fish Contamination Program.
Tudor T. Davies, Director
Office of Science and Technology
Recycled/Recyclable .Printed with Vegetable Oil Based Inks on 100% Recycled Paper (40% Postconsumer)
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TABLE OF CONTENTS
TABLE OF CONTENTS
Section
Page
*
List of Figures xiv
List of Tables xv
Glossary and Abbreviations xix
Acknowledgments xxv
Executive Summary xxviii
Introduction 1-1
1.1 Overview and Objectives 1-1
1.2 Contents of Volume 2 1-3
1.3 Sources 1-7
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
2.2.1.1 Toxicological Data 2-5
2.2.1.2 Contaminant Data 2-6
2.2.1.3 Sources of Exposure 2-7
2.2.2 Assumptions and Uncertainty Analysis 2-8
2.3 Dose-Response Assessment 2-9
2.3.1 Carcinogenic Effects 2-11
2.3.2 Noncarcinogenic Effects ; 2-13
2.3.2.1 Acute Exposure 2-13
2.3.2.2 Systemic Effects from Chronic Exposure 2-14
2.3.2.3 Developmental Toxicity 2-18
2.3.3 Mutagenicity/Genotoxicity 2-28
2.3.4 Multiple Chemical Exposures: Interactive Effects 2-28
2.3.5 Assumptions and Uncertainties 2-30
2.4 Exposure Assessment 2-33
2.4.1 Chemical Occurrences in Fish 2-34
2.4.2 Geographic Distribution of Contaminated Fish 2-35
2.4.3 Individual Exposure Assessment 2-35
2.4.3.1 Exposure Variables 2-36
2.4.3.2 Averaging Periods Versus Exposure
Durations 2-39
2.4.3.3 Multiple Species Exposures 2-41
2.4.3.4 Multiple Chemical Exposures 2-42
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TABLE OF CONTENTS
Section
Page
2.4.4 Population Exposure Assessments ., 2-42
2.4.4.1 Categories of Population Exposure
Assessment Information 2-42
2.4.4.2 Categorizing Exposure Levels 2-45
2.4.5 Uncertainty and Assumptions 2-48
2.4.5.1 Chemical Contaminant Concentrations
in Fish • • 2-48
2.4.5.2 Dose Modifications Due to Food Preparation
and Cooking 2-49
2.4.5.3 Body Weight 2-49
2.4.5.4 Consumption Rate and Averaging Period 2-49
2.4.5.5 Multiple Species and Multiple Contaminants .... 2-52
2.4.5.6 Other Sources of Exposure 2-53
2.5 Risk Characterization 2-58
2.5.1 Carcinogenic Toxicity 2-59
2.5.1.1 Individual Risk 2-60
2.5.1.2 Population Risk , 2-60
2.5.2 Noncarcinogenic Toxicity 2-61
2.5.2.1 Individual Risk 2-61
2.5.2.2 Population Risk 2-61
2.5.3 Subpopulation Considerations 2-62
2.5.4 Multiple Species and Multiple Contaminant
Considerations 2-64
2.5.5 Incorporating Considerations of Uncertainty in
Consumption Limits 2-64
2.6 Summarizing Risk Data 2-65
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
3.2.1.1 Calculation of Daily Consumption Limits 3-5
3.2.1.2 Calculation of Meal Consumption Limits 3-5
3.2.1.3 Input Parameters 3-6
3.2.2 Calculation of Consumption Limits for
Noncarcinogenic Effects 3-6
3.2.2.1 Calculation of Daily Consumption Limits .3-9
3.2.2.2 Calculation of Meal Consumption Limits 3-11
3.2.2.3 Input Parameters 3-12
3.2.3 Calculation of Consumption Limits for Developmental
Effects '• 3-12
3.3 Default and Alternative Values for Calculating Consumption
Limits 3-15
3.3.1 Maximum Acceptable Risk Level 3-15
iv
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TABLE OF CONTENTS
Section
Page
3.3.2 Cancer Potencies and Chronic Reference Doses
(q.,*s and RfDs) . 3-13
3.3.3 Consumer Body Weight (BW) 3-20
3.3.3.1 Derivation of Multipliers for Body Weight
Adjustment .'... 3-21
3.3.4 Meal Size 3-23
3.3.5 Contaminant Concentration in Fish Tissue 3-28
3.3,6 Modifying Time-Averaging Period (Tap) 3-29
3.4 Modification of Consumption Limits for a
Single Contaminant in a Multispecies Diet 3-31
3.4.1 Carcinogenic Effects 3-32
3.4.2 Noncarcinogenic Effects 3.33
3.5 Modification of Consumption Limits for
Multiple Contaminant Exposures 3-35
3.5.1 Carcinogenic Effects 3-36
3.5.2 Noncarcinogenic Effects 3-37
3.5.3 Species-Specific Consumption Limits in a Multiple
Species Diet , 3.39
3.6 Choice of Consumption Limits 3-40
Risk-Based Consumption Limit Tables 4-1
4.1 Overview and Section Organization 4-1
4.2 Consumption Limit Tables 4.4
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
5.1.1.1 Pharmacokinetics 5-2
5,1.1.2 Acute Toxicity 5-4
5.1.1.3 Chronic Toxicity 5.4
5.1.1.4 Developmental Toxicity 5-4
5.1.1.5 Mutagenicity 5-5
5.1.1.6 Carcinogenicity 5-5
5.1.1.7 Special Susceptibilities 5-6
5.1.1.8 Interactive Effects 5-7
5.1.11.9 Critical Data Gaps 5-7
5.1.1.10 Summary of EPA Levels of Concern 5-7
5.1.1.11 Major Sources 5-7
5.1.1.12 Statement Regarding Uncertainty 5-8
5.1.2 Abbreviations Used and Scientific Notation 5-8
5.2 Metals 5-10
5.2.1 Arsenic 5-10
5.2.1.1 Background 5-10
5.2.1.2 Pharmacokinetics 5-10
5.2.1.3 Acute Toxicity 5-10
5.2.1.4 Chronic Toxicity 5-1 o
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TABLE OF CONTENTS
Section
Page
5.2.1.5 Developmental Toxicity 5-11
5.2.1.6 Mutagenicity 5-12
5.2.1.7 Carcinogenicity 5-12
5.2.1.8 SpecialSusceptibilities 5-12
5.2.1.9 Interactive Effects 5-13
5.2.1.10 Critical Data Gaps 5-13
5.2.1.11 Summary of EPA Levels of Concern 5-13
5.2.1.12 MajorSources 5-13
5.2.2 Cadmium 5-13
5.2.2.1 Background 5-13
5.2.2.2 Pharmacokinetics 5-14
5.2.2.3 Acute Toxicity 5-14
5.2.2.4 Chronic Toxicity 5-15
5.2.2.5 Developmental Toxicity 5-16
5.2.2.6 Mutagenicity 5-17
5.2.2.7 Carcinogenicity 5-17
5.2.2.8 Special Susceptibilities 5-18
5.2.2.9 Interactive Effects 5-18
5.2.2.10 Critical Data Gaps 5-18
5.2.2.11 Summary of EPA Levels of Concern 5-19
5.2.2.12 Major Sources 5-19
5.2.3 Mercury 5-19
5.2.3.1 Background 5-19
5.2.3.2 Pharmacokinetics 5-20
5.2.3.3 Acute Toxicity 5-20
5.2.3.4 Chronic Toxicity 5-20
5.2.3.5 Developmental Toxicity 5-24
5.2.3.6 Mutagenicity , 5-25
5.2.3.7 Carcinogenicity 5-25
5.2.3.8 Special Susceptibilities 5-25
5.2.3.9 Interactive Effects 5-26
5.2.3.10 Critical Data Gaps ' 5-26
5.2.3.11 Summary of EPA Levels of Concern 5-26
5.2.3.12 Major Sources 5-26
5.2.4 Selenium 5-26
5.2.4.1 Background 5-26
5.2.4.2 Pharmacokinetics 5-27
5.2.4.3 Acute Toxicity 5-27
5.2.4.4 Chronic Toxicity 5-27
5.2.4.5 Developmental Toxicity 5-29
5.2.4.6 Mutagenicity 5-31
5.2.4.7 Carcinogenicity 5-31
5.2.4.8 Special Susceptibilities 5-31
5.2.4.9 Interactive Effects 5-32
5.4.2.10 Critical Data Gaps 5-32
5.2.4.11 Summary of EPA Levels of Concern 5-32
vi
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TABLE OF CONTENTS
Section
Page
5.2.4.12 Major Sources 5-32
5.2.5 Tributyltin Oxide 5.32
5.2.5.1 Background '. 5-32
5.2.5.2 Pharmacokinetics 5-33
5.2.5.3 Acute Toxicity 5-33
5.2.5.4 Chronic Toxicity 5-33
5.2.5.5 Developmental Toxicity 5-34
5.2.5.6 Mutagenicity 5-34
5.2.5.7 Carcinogenicity 5-34
5.2.5.8 Special Susceptibilities 5-34
5.2.5.9 Interactive Effects 5-35
5.2.5.10 Critical Data Gaps 5-35
5.2.5.11 Summary of EPA Levels of Concern 5-35
5.2.5.12 Major Sources 5-35
5.3 Organochlorine Pesticides 5-36
5.3.1 Chlordane 5-36
5.3.1.1 Background 5-36
5.3.1.2 Pharmacokinetics 5-35
5.3.1.3 Acute Toxicity .... '. 5-35
5.3.1.4 Chronic Toxicity 5-37
5.3.1.5 Developmental Toxicity 5-37
5.3.1.6 Mutagenicity 5-33
5.3.1.7 Carcinogenicity 5-33
5.3.1.8 Special Susceptibilities 5-39
5.3.1.9 Interactive Effects 5-39
5.3.1.10 Critical Data Gaps 5-40
5.3.1.11 Summary of EPA Levels of Concern 5-40
5.3.1.12 Major Sources 5-40
5.3.2 DDT, DDE, ODD 5-40
5.3.2.1 Background 5-40
5.3.2.2 Pharmacokinetics 5-41
5.3.2.3 Acute Toxicity 5-41
5.3.2.4 Chronic Toxicity 5-41
5.3.2.5 Developmental Toxicity 5-42
5.3.2.6 Mutagenicity 5-44
5.3.2.7 Carcinogenicity 5-44
5.3.2.8 Special Susceptibilities 5-44
5.3.2.9 Interactive Effects 5-45
5.3.2.10 Critical Data Gaps 5.45
5.3.2.11 Summary of EPA Levels of Concern 5-46
5.3.2.12 Major Sources .: 5-45
5.3.3 Dicofol (Kelthane) 5.45
5.3.3.1 Background 5-46
5.3.3.2 Pharmacokinetics 5.45
5.3.3.3 Acute Toxicity 5.45
5.3.3.4 Chronic Toxicity 5-45
vii
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TABLE OF CONTENTS
Section
Page
5.3.3.5 Developmental Toxicity 5-47
5.3.3.6 Mutagenicity 5-47
5.3.3.7 Carcinogenicity 5-48
5.3.3.8 Special Susceptibilities 5-48
5.3.3.9 Interactive Effects 5-48
5.3.3.10 Critical Data Gaps 5-48
5.3.3.11 Summary of EPA Levels of Concern 5-48
5.3.3.12 Major Sources • 5-48
5.3.4 Dieldrin 5-49
5.3.4.1 Background 5-49
5.3.4.2 Pharmacokinetics 5-49
3.3.4.3 Acute Toxicity 5-49
5.3.4.4 Chronic Toxicity 5-49
5.3.4.5 Developmental Toxicity 5-51
5.3.4.6 Mutagenicity 5-53
5.3.4.7 Carcinogenicity 5-53
5.3.4.8 Special Susceptibilities 5-54
5.3.4.9 Interactive Effects 5-54
5.3.4.10 Critical Data Gaps 5-54
5.3.4.11 Summary of EPA Levels of Concern 5-54
5.3.4.12 Major Sources 5-54
5.3.5 Endosulfan I, II 5-55
5.3.5.1 Background • 5-55
5.3.5.2 Pharmacokinetics 5-55
5.3.5.3 Acute Toxicity 5-55
5.3.5.4 Chronic Toxicity 5-55
5.3.5.5 Developmental Toxicity 5-56
5.3.5.6 Mutagenicity 5-57
5.3.5.7 Carcinogenicity 5-57
5.3.5.8 Special Susceptibilities 5-58
5.3.5.9 Interactive Effects 5-58
5.3.5.10 Critical Data Gaps 5-58
5.3.5.11 Summary of EPA Levels of Concern 5-59
5.3.5.12 Major Sources 5-59
5.3.6 Endrin 5-59
5.3.6.1 Background 5-59
5.3.6.2 Pharmacokinetics 5-59
5.3.6.3 Acute Toxicity 5-59
5.3.6.4 Chronic Toxicity 5-59
5.3.6.5 Developmental Toxicity 5-60
5.3.6.6 Mutagenicity 5-61
5.3.6.7 Carcinogenicity 5-61
5.3.6.8 Special Susceptibilities 5-61
5.3.6.9 Interactive Effects 5-61
5.3.6.10 Critical Data Gaps 5-62
5.3.6.11 Summary of EPA Levels of Concern 5-62
VIM
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TABLE OF CONTENTS
Section
Page
5.3.6.12 Major Sources 5-62
5.3.7 Heptachlor Epoxide 5-62
5.3.7.1 Background 5-62
5.3.7.2 Pharmacokinetics 5-62
5.3.7.3 Acute Toxicity 5-63
5.3.7.4 Chronic Toxicity 5-63
5.3.7.5 Developmental Toxicity 5-63
5.3.7.6 Mutagenicity 5-64
5.3.7.7 Carcinogenicity 5-64
5.3.7.8 Special Susceptibilities 5-65
5.3.7.9 Interactive Effects 5-65
5.3.7.10 Critical Data Gaps 5-65
5.3.7.11 Summary of EPA Levels of Concern 5-65
5.3.7.12 Major Sources 5-65
5.3.8 Hexachlorobenzene 5-65
5.3.8.1 Background 5-65
5.3.8.2 Pharmacokinetics 5-66
5.3.8.3 Acute Exposure 5-66
5.3.8.4 Chronic Toxicity 5-66
5.3.8.5 Developmental Toxicity 5-67
5.3.8.6 Mutagenicity 5-68
5.3.8.7 Carcinogenicity 5-68
5.3.8.8 Special Susceptibilities 5-68
5.3.8.9 Interactive Effects 5-69
5.3.8.10 Critical Data Gaps 5-69
5.3.8.11 Summary of EPA Levels of Concern 5-69
5.3.8.12 Major Sources 5-69
5.3.9 Lindane (v-hexachlorocyclohexane) 5-69
5.3.9.1 Background 5-69
5.3.9.2 Pharmacokinetics 5-69
5.3.9.3 Acute Toxicity 5-70
5.3.9.4 Chronic Toxicity 5-70
5.3.9.5 Developmental Toxicity 5-71
5.3.9.6 Mutagenicity 5-72
5.3.9.7 Carcinogenicity 5-72
5.3.9.8 Special Susceptibilities 5-73
5.3.9.9 Interactive Effects 5-73
5.3.9.10 Critical Data Gaps 5-73
5.3.9.11 Summary of EPA Levels of Concern 5-74
5.3.9.12 Major Sources 5-74
5.3.10 Mirex 5-74
5.3.10.1 Background : 5-74
5.3.10.2 Pharmacokinetics 5-74
5.3.10.3 Acute Toxicity 5-75
5.3.10.4 Chronic Toxicity 5-75
5.3.10.5 Developmental Toxicity 5-76
IX
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TABLE OF CONTENTS
Section
Page
5.3.10.6 Mutagenicity 5-77
5.3.10.7 Carcinogenicity 5-77
5.3.10.8 Special Susceptibilities 5-77
5.3.10.9 Interactive Effects 5-78
5.3.10.10 Critical Data Gaps 5-78
5.3.10.11 Summary of EPA Levels of Concern 5-78
5.3.10.12 Major Sources 5-78
5.3.11 Toxaphene 5-78
5.3.11.1 Background 5-78
5.3.11.2 Pharmacokinetics 5-78
5.3.11.3 Acute Toxicity 5-79
5.3.11.4 Chronic Toxicity 5-79
5.3.11.5 Developmental Toxicity 5-79
5.3.11.6 Mutagenicity 5-81
5.3.11.7 Carcinogenicity 5-81
5.3.11.8 Special Susceptibilities 5-81
5.3.11.9 Interactive Effects 5-82
5.3.11.10 Critical Data Gaps 5-83
5.3.11.11 Summary of EPA Levels of Concern 5-83
5.3.11.12 Major Sources 5-83
5.4 Organophosphate Pesticides 5-84
5.4.1 Chlorpyrifos 5-84
5.4.1.1 Background 5-84
5.4.1.2 Pharmacokinetics 5-84
5.4.1.3 Acute Toxicity 5-84
5.4.1.4 Chronic Toxicity 5-84
5.4.1.5 Developmental Toxicity 5-85
5.4.1.6 Mutagenicity 5-85
5.4.1.7 Carcinogenicity 5-86
5.4.1.8 Special Susceptibilities 5-86
5.4.1.9 Interactive Effects 5-86
5.4.1.10 Critical Data Gaps 5-86
5.4.1.11 Summary of Levels of Concern 5-86
5.4.1.12 Major Sources 5-86
5.4.2 Diazinon 5-86
5.4.2.1 Background 5-86
5.4.2.2 Pharmacokinetics 5-86
5.4.2.3 Acute Toxicity 5-87
5.4.2.4 Chronic Toxicity 5-87
5.4.2.5 Developmental Toxicity 5-87
5.4.2.6 Mutagenicity 5-88
5.4.2.7 Carcinogenicity 5-88
5.4.2.8 Special Susceptibilities 5-88
5.4.2.9 Interactive Effects 5-88
5.4.2.10 Critical Data Gaps 5-88
5.4.2.11 Summary of EPA Levels of Concern 5-88
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TABLE OF CONTENTS
Section
Page
5.4.2.12 Major Sources 5-88
5.4.3 Disulfoton (disyston) 5-88
5.4.3.1 Background 5-88
5.4.3.2 Pharmacokinetics 5-89
5.4.3.3 Acute Toxicity 5-89
5.4.3.4 Chronic Toxicity 5-89
5.4.3.5 Developmental Toxicity 5-89
5.4.3.6 Mutagenicity 5-90
5.4.3.7 Carcinogenicity 5-90
5.4.3.8 Special Susceptibilities 5-90
5.4.3.9 Interactive Effects 5-91
5.4.3.10 Critical Data Gaps 5-91
5.4.3.11 Summary of EPA Levels of Concern 5-91
5.4.3.12 Major Sources 5-91
5.4.4 Ethion 5-91
5.4.4.1 Background 5-91
5.4.4.2 Pharmacokinetics 5-91
5.4.4.3 Acute Toxicity 5-91
5.4.4.4 Chronic Toxicity 5-91
5.4.4.5 Developmental Toxicity 5-92
5.4.4.6 Mutagenicity 5-92
5.4.4.7 Carcinogenicity ., 5-92
5.4.4.8 Special Susceptibilities 5-92
5.4.4.9 Interactive Effects 5-92
5.4.4.10 Critical Data Gaps 5-93
5.4.4.11 Summary of EPA Levels of Concern 5-93
5.4.4.12 Major Sources 5-93
5.4.5 Terbufos 5-93
5.4.5.1 Background 5-93
5A5.2 Pharmacokinetics 5-93
5.4.5.3 Acute Toxicity 5-93
5.4.5.4 Chronic Toxicity ; 5-93
5.4.5.5 Developmental Toxicity 5-94
5.4.5.6 Mutagenicity 5-95
5.4.5.7 Carcinogenicity 5-95
5.4.5.8 Special Susceptibilities 5-95
5.4.5.9 Interactive Effects 5-95
5.4.5.10 Critical Data Gaps 5-95
5.4.5.11 Summary of EPA Levels of Concern 5-96
5.4.5;12 Major Sources 5-96
5.5 Chlorophenoxy Herbicides , 5-97
5.5.1 Oxyfluorfen 5-97
5.5.1.1 Background 5-97
5.5.1.2 Pharmacokinetics 5-97
5.5.1.3 Acute Toxicity 5-97
5.5.1.4 Chronic Toxicity 5-97
XI
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TABLE OF CONTENTS
Section
Page
5.5.1.5 Developmental Toxicity 5-97
5.5.1.6 Mutagenicity 5-97
5.5.1.7 Carcinogenicity 5-98
5.5.1.8 Interactive Effects 5-98
5.5.1.9 Critical Data Gaps 5-98
5.5.10 Summary of EPA Levels of Concern 5-98
5.5.1.11 Major Sources 5-98
5.6 Polycyclic Aromatic Hydrocarbons (PAHs) 5-99
5.6.1 Background 5-99
5.6.2 Pharmokinetics 5-99
5.6.3 Acute Toxicity 5-100
5.6.4 Chronic Toxicity 5-100
5.6.5 Developmental Toxicity 5-100
5.6.6 Mutagenicity 5-100
5.6.7 Carcinogenicity 5-101
5.6.8 Special Susceptibilities 5-102
5.6.9 Interactive Effects 5-103
5.6.10 Critical Data Gaps 5-104
5.6.11 Summary of EPA Levels of Concern '... 5-104
5.6.12 Major Sources 5-104
5.7 Polychlorinated Biphenyls (PGBs) 5-105
5.7.1 Background 5-105
5.7.2 Pharmacokinetics 5-105
5.7.3 Acute Toxicity 5-106
5.7.4 Chronic Toxicity 5-107
5.7.5 Developmental Toxicity 5-107
5.7.6 Mutagenicity 5-111
5.7.7 Carcinogenicity 5-111
5.7.8 Special Susceptibilities 5-112
5.7.9 Interactive Effects 5-113
5.7.10 Critical Data Gaps 5-113
5.7.11 Summary of EPA Levels of Concern 5-113
5.7.12 Major Sources 5-113
5.8 Dioxins 5-114
5.8.1 Background 5-114
5.8.2 Summary of EPA Levels of Concern 5-114
5.8.3 Major Source 5-114
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 6-8
xii
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TABLE OF CONTENTS
Section
Appendix
A
B
C
D
E
F
G
H
Page
Literature Cited 7-1
Sources of Additional Information A-1
Mutagenicity and Genotoxicity Guidelines B-1
Toxicity Characteristics of Groups of Analytes C-1
Population Exposure Assessment—Consumption Patterns and
Surveys D-1
Dose Modification Due to Food Preparation and Cooking E-1
Guidance for Risk Characterization F-1
Sample Consumption Limit Tables G-1
Summary of Limits of Detection for the Recommended
Target Analytes H-1
XIII
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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 ; 1-5
2-1 Elements of Risk Assessment and Risk Management 2-2
2-2 Schematic of Exposure Categories in Upper Half of a Normal
Population Distribution 2-46
3-1 Current RfD for Methylmercury with Associated Range of Uncertainty and
Corresponding Fish Tissue Contamination Levels 3-10
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 , 6-7
6-3 Map Showing Active Fish and Wildlife Advisories for a State 6-7
XIV
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LIST OF TABLES
LIST OF TABLES
Number Page
1-1 Target Analytes Recommended for Fish Sampling Programs 1-4
2-1 Uncertal nty Factors and Modifying Factors for Estimating Exposure Limits for
Developmental Effects 2-16
2-2 Mean Body Weights of Children and Adults 2-37
2-3 Categories of Information Necessary for a Population
Exposu -e Assessment 2-43
2-4 Exposu-e Data Template 2-56
2-5 Risk Estimates 2-66
2-6 Risk Cr aracterization 2-67
2-7 Risk Summaries for a Waterbody 2-68
2-8 Risk Summaries for a Geographic Area 2-70
3-1 Risk Vaj ues Used in Risk-Based Consumption Limit Tables 3-3
3-2 Input Parameters for Use in Risk Equations 3-7
3-3 Monthly Consumption Limits for Carcinogenic Health Endpoints
for the General Population—Chlordane 3-8
3-4 Monthly] Consumption Limits for Chronic Systemic Health Endpoints
for the General Population—Chlordane 3-13
3-5 Monthly Consumption Limits for Developmental Health Endpoints
for Women of Reproductive Age—PCBs (Aroclor 1016) 3-16
3-6 Monthly] Consumption Limits for Chronic Systemic Health Endpoints
for the General Population—PCBs (Aroclor 1254) 3-17
3-7 Monthly] Consumption Limits for Carcinogenic Health Endpoints
for the General Population—Chlordane 3-19
3-8 Average Body Weights and Associated Multipliers 3-21
3-9 Monthly Consumption Limits for Chronic Systemic Health Endpoints
for Children—Chlordane 3-24
3-10 Alternalive Meal Sizes and Associated Multipliers 3-25
3-11 Monthly Consumption Limits for Chronic Systemic Health Endpoints
for the General Population—Chlordane 3-27
3-12 Monthly Consumption Limits for Chronic Systemic Health Endpoints
for Children—Chlordane , 3-29
3-13 10-Day Consumption Limits for Chronic Systemic Health Endpoints
for the General Population—Chlordane 3-32
xv
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LIST OF TABLES
Number
Page
4-1 Index to Monthly Consumption Limit Tables for Target Analytes 4-5
4-2 Summary of RfD, CSF, and Detection Limit Values Used to Develop
Fish Consumption Limits 4-6
4-3 Monthly Consumption Limits for Chronic Systemic Health Endpoints
for the General Population—Arsenic (inorganic) 4-7
4-4 Monthly Consumption Limits for Chronic Systemic Health Endpoints for
Children—Arsenic (inorganic) 4-8
4-5 Monthly Consumption Limits for Carcinogenic Health Endpoints for
the General Population—Arsenic (inorganic) 4-9
4-6 Monthly Consumption Limits for Chronic Systemic Health Endpoints for
the General Population—Cadmium 4-10
4-7 Monthly Consumption Limits for Chronic Systemic Health Endpoints for
Children—Cadmium 4-11
4-8 Monthly Consumption Limits for Chronic Systemic Health Endpoints for :
the General Population—Methylmercury (IRIS RfD) 4-12
4-9 Monthly Consumption Limits for Developmental Health Endpoints for Women
of Reproductive Age—Methylmercury (IRIS RfD) 4-13
4-10 Monthly Consumption Limits for Developmental Health Endpoints for
Children—Methylmercury (IRIS RfD) 4-14
4-11 Monthly Consumption Limits for Chronic Systemic Health Endpoints
for the General Population—Selenium 4-15
4-12 Monthly Consumption Limits for Chronic Systemic Health Endpoints
for Children—Selenium 4-16
4-13 Monthly Consumption Limits for Chronic Systemic Health Endpoints
for the General Population—Tributyltin Oxide 4-17
4-14 Monthly Consumption Limits for Chronic Systemic Health Endpoints
for Children—Tributyltin Oxide 4-18
4-15 Monthly Consumption Limits for Chronic Systemic Health Endpoints
for the General Population—Chlordane 4-19
4-16 Monthly Consumption Limits for Chronic Systemic Health Endpoints
for Children—Chlordane 4-20
4-17 Monthly Consumption Limits for Carcinogenic Health Endpoints
for the General Population—Chlordane 4-21
4-18 Monthly Consumption Limits for Chronic Systemic Health Endpoints
forthe General Population—DDT 4-22
4-19 Monthly Consumption Limits for Chronic Systemic Health Endpoints
for Children—DDT 4-23
4-20 Monthly Consumption Limits for Carcinogenic Health Endpoints
for the General Population—DDT 4-24
4-21 Monthly Consumption Limits for Chronic Systemic Health Endpoints
for the General Population—Dicofol 4-25
4-22 Monthly Consumption Limits for Chronic Systemic Health Endpoints
for Children—Dicofol 4-26
4-23 Monthly Consumption Limits for Carcinogenic Health Endpoints
for the General Population—Dicofol 4-27
xvi
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LIST OF TABLES
Number
Page
4-24 Monthly Consumption Limits for Chronic Systemic Health Endpoints
for the General Population—Dieldrin 4-28
4-25 Monthly Consumption Limits for Chronic Systemic Health Endpoints
for Children—Dieldrin 4-29
4-26 Monthly Consumption Limits for Carcinogenic Health Endpoints
for the General Population^Dieldrin 4-30
4-27 Monthly Consumption Limits for Chronic Systemic Health Endpoints
forthe General Population—Endosulfan 4-31
4-28 Monthly Consumption Limits for Chronic Systemic Health Endpoints
for Children—Endosulfan ;...................... 4-32
4-29 Monthly Consumption Limits for Chronic Systemic Health Endpoints
for the General Population—Endrin 4-33
4-30 Monthly Consumption Limits for Chronic Systemic Health Endpoints
for Children—Endrin 4-34
4-31 Monthly Consumption Limits for Chronic Systemic Health Endpoints
for the General Population—Heptachlor Epoxide 4-35
4-32 Monthly Consumption Limits for Chronic Systemic Health Endpoints
for Children—Heptachlor Epoxide 4-36
4-33 Monthly Consumption Limits for Carcinogenic Health Endpoints
for the General Population—Heptachlor Epoxide 4-37
4-34 Monthly Consumption Limits for Chronic Systemic Health Endpoints
forthe General Population—Hexachlorobenzene 4-38
4-35 Monthly Consumption Limits for Chronic Systemic Health Endpoints
for Children—Hexachlorobenzene 4-39
4-36 Monthly Consumption Limits for Carcinogenic Health Endpoints
for the General Population—Hexachlorobenzene 4-40
4-37 Monthly Consumption Limits for Chronic Systemic Health Endpoints
for the General Population—Lindane 4-41
4-38 Monthly Consumption Limits for Chronic Systemic Health Endpoints
for Children—Lindane 4-42
4-39 Monthly Consumption Limits for Carcinogenic Health Endpoints
for the General Population—Lindane 4-43
4-40 Monthly Consumption Limits for Chronic Systemic Health Endpoints
for the General Population—Mirex 4-44
4-41 Monthly Consumption Limits for Chronic Systemic Health Endpoints
for Children—Mirex 4-45
4-42 Monthly Consumption Limits for Carcinogenic Health Endpoints
for the General Population—Mirex 4-46
4-43 Monthly Consumption Limits for Chronic Systemic Health Endpoints
for the General Population—Toxaphene (OW/OPP RfD) 4-47
4-44 Monthly Consumption Limits for Chronic Systemic Health Endpoints
for Children—Toxaphene (OW/OPP RfD) 4-48
4-45 Monthly Consumption Limits for Carcinogenic Health Endpoints
for the General Population—Toxaphene , 4-49
4-46 Monthly Consumption Limits for Chronic Systemic Health Endpoints
for the General Population—Chlorpyrifos 4-50
xvii
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LIST OF TABLES
Number
Page
4-47 Monthly Consumption Limits for Chronic Systemic Health Endpoints
for Children—Chlorpyrifos 4-51
4-48 Monthly Consumption Limits for Chronic Systemic Health Endpoints
forthe General Population—Diazinon 4-52
4-49 Monthly Consumption Limits for Chronic Systemic Health Endpoints
for Children—Diazinon 4-53
4-50 Monthly Consumption Limits for Chronic Systemic Health Endpoints
for the General Population—Disulfoton 4-54
4-51 Monthly Consumption Limits for Chronic Systemic Health Endpoints
for Children—Disulfoton 4-55
4-52 Monthly Consumption Limits for Chronic Systemic Health Endpoints
for the General Population—Ethion 4-56
4-53 Monthly Consumption Limits for Chronic Systemic Health Endpoints
for Children—Ethion 4-57
4-54 Monthly Consumption Limits for Chronic Systemic Health Endpoints
for the General Population—Terbufos 4-58
4-55 Monthly Consumption Limits for Chronic Systemic Health Endpoints
for Children—Terbufos 4-59
4-56 Monthly Consumption Limits for Chronic Systemic Health Endpoints
for the General Population—Oxyfluorfen 4-60
4-57 Monthly Consumption Limits for Chronic Systemic Health Endpoints
for Children—Oxyfluorfen 4-61
4-58 Monthly Consumption Limits for Carcinogenic Health Endpoints
forthe General Population—Oxyfluorfen 4-62
4-59 Monthly Consumption Limits for Carcinogenic Health Endpoints
forthe General Population—PAHs 4-63
4-60 Monthly Consumption Limits for Chronic Systemic Health Endpoints
forthe General Population—PCBs (Arochlor 1254) 4-64
4-61 Monthly Consumption Limits for Developmental Health Endpoints
for Women of Reproductive Age—PCBs (Arochlor 1254) 4-65
4-62 Monthly Consumption Limits for Developmental Health Endpoints
for Children—PCBs (Arochlor 1254) 4-66
4-63 Monthly Consumption Limits for Carcinogenic Health Endpoints
forthe General Population—PCBs (Total) 4-67
4-64 Monthly Consumption Limits for Carcinogenic Health Endpoints
for the General Population—Dioxins/Furans 4-68
5-1 Health and Toxicological Data Reviewed for Target Analytes 5-3
5-2 Toxicity Equivalent Factors for Various PAHs 5-102
5-3 Relative Potency Estimates for Various PAHs 5-103
6-1
Comparison of Raster- Versus Vector-Based GIS Programs 6-4
XVIII
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GLOSSARY AND ABBREVIATIONS
GLOSSARY AND ABBREVIATIONS
acute exposure
ARL
ATSDR
BW
GAG
cancer slope factor
carcinogen
CCRIS
CERCLA
CERCLIS
chronic exposure
cm
CNS
exposure at a relatively high level over a short period of time
(minutes to a few days) (This is defined in IRIS.as 24 hours or
less; however, sources consulted utilized exposure periods of up
to a few days. Consequently, the more encompassing definition is
appropriate in reading this document.)
acceptable risk level. The maximum level of individual lifetime
carcinogenic risk, usually calculated using a cancer slope value,
which is considered "acceptable" by risk managers. See Section 4
for more detail.
Agency for Toxic Substances and Disease Registry, U.S.
Department of Health and Human Services, Public Health Service
body weight of an individual consumer (kg)
Career Assessment Group at U.S. Environmental Protection
Agency
the slope of the dose-response curve in the low-dose region used
with exposure to calculate the estimated lifetime cancer risk. Often
expressed as risk per 1 m of exposure to the toxic chemical per
kilogram body weight per day (mg/kg/d).
an agent capable of inducing a carcinogenic response
Chemical Carcinogenesis Research Information System
Comprehensive Environmental Response, Compensation, and
Liability Act
Comprehensive Environmental Response, Compensation, and
Liability Act List of Sites
multiple exposures occurring over an extended period of time, or a
significant fraction of the lifetime
concentration of contaminant m in the edible portion of fish
species j (mg/kg)
concentration of contaminant m in the edible portion of fish
(mg/kg)
central nervous system
XIX
-------
GLOSSARY AND ABBREVIATIONS
confounder
consumption limits
OR
critical effect
OR
'md
OR
'mm
'tin
developmental toxicity
dose-response
mj
me
endpoint
EPA
exposure limits
FDA
PEL
FGDC
condition or variable that may be a factor in producing the same
response as the agent under study
a daily fish consumption limit, based on health and toxicity data
mean daily consumption rate of fish (kg/d)
the first adverse effect, or its known precursor, that occurs as the
dose rate increases
consumption rate of fish species / (kg/d)
limit on the amount of fish that can be consumed per day (kg/d)
limit on the number of fish meals that can be consumed per day
(meals/day)
limit on the number of fish meals that can be consumed per month
(meals/mo)
total concentration of chemical m in an individual's fish diet
study of adverse effects on the developing organism resulting from
exposure prior to conception, during prenatal development, or
postnatally up to the time of sexual maturation
relationship between the amount of an agent and changes in
aspects of the biological system apparently in response to that
agent
exposure to contaminant m from ingesting fish (mg/kg/d)
exposure to contaminant m from ingesting fish species / (mg/kg/d)
exposure to a given contaminant in a given species of fish
associated with a given risk of cancer (mg/kg/d)
maximum acceptable exposure (dose) of a noncarcinogen from a
specific contaminant in a specific fish species (mg/kg/d)
response measure in a toxicity study
U.S. Environmental Protection Agency
a daily limit on exposure based on health and toxicity data, which
the reader may calculate, using the study data provided in this or
other sources (mg/kg/d)
U.S. Food and Drug Administration
frank effect level
Federal Geographic Data Committee
xx
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GLOSSARY AND ABBREVIATIONS
FRAG
GIS
GPS
HCG
HEAST
HI
H'mix
HSDB
incidence
IRIS
latency period
LEL
LH
LMS
LOAEL
LOD
modifying factor
MOE
MRL
MS
mutagenic
NAFWCD
fraction of a given fish species in an individual's diet (unitless)
geographic information system
Global Positioning Satellite
human chorionic gonadotropic
Health Effects Assessment Summary Tables
hazard index, or ratio of the estimated exposure dose to the RfD
for the chemical (unitless)
hazard index of a chemical mixture (unitless)
Hazardous Substances Data Bank, available on line through
TOXNET
number of new cases of a disease within a specified time
Integrated Risk Information System, a database maintained by
EPA, available on line through TOXNET and by subscription
through NTIS
time between induction of a health effect and its manifestation
same as LOAEL (per IRIS), see below
luteinizing hormone
linearized multistage model
lowest exposure level at which there are statistically or biologically
significant increases in frequency of severity of adverse effects
between the exposed population and its appropriate control group
limit of detection
a factor used in operationally deriving the RfD from experimental
data. It addresses concerns regarding differences in absorption,
tolerance to a chemical, or lack of a sensitive endpoint.
margin of exposure
Minimal Risk Level, from ATSDR. An estimate of daily exposure
that is likely to be without an appreciable risk of deleterious effects
(noncancerous) over a specified duration of exposure: acute—1 to
13 days; intermediate—14 to 365 days; chronic—over 365 days
meal size (kg/meal)
capable of inducing changes in genetic material (e.g., DNA)
North American Fish and Wildlife Consumption Database
XXI
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GLOSSARY AND ABBREVIATIONS
NAS
NGOs
NHANES II
NIOSH
MOM
NOAEL
J1
NOEL
NSDI
NTIS
NTP
OAPCA
OHEA
OPP
PAHS
PCBs
PCS
PEC
PJ
PNA
PNS
POTW
31*
QA/QC
R
RAG
National Academy of Sciences
nongovernmental organizations
National Health and Nutrition Examination Survey
National Institute of Occupational Safety and Health
National Oceanic and Atmospheric Association
exposure level at which there are no statistically or biologically
significant increases in the frequency or severity of adverse effects
between the exposed population and its control
the same as NOAEL with the exception of the word adverse.
NOEL specifies the absence of any effect
National Spatial Data Infrastructure
National Technical Information Service
National Toxicology Program
Organotin Antifouling Paint Control Act
Office of Health Effects Assessment, U.S. Environmental
Protection Agency
Office of Pesticide Programs, U.S. Environmental Protection
Agency
polyaromatic hydrocarbons
polychlorinated biphenyls
Permit Compliance Systems
potency equivalency concentration
proportion of a given species in the diet (unitless)
polynuclear aromatic hydrocarbon
peripheral nervous system
publicly owned treatment works
cancer slope factor, lifetime cancer risk per mg/kg/d
quality assurance/quality control
incremental risk above background associated with contaminant
at given dose (lifetime"1)
reference ambient concentrations
xxii
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GLOSSARY AND ABBREVIATIONS
Ravg
RDA
reference dose (RfD)
RfD
risk
m
RL
Rmix
RSC
RTECS
SAB
SCE
SF
SFm
SIZ
svc
svn
TCDD
TEC
TEF
teratogenic
threshold
TRI
mean individual risk in the exposed population (risk/person-
lifetime)
Recommended Dietary Allowance
estimate (with uncertainty spanning perhaps 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 deleterious effects during a lifetime (mg/kg/d).
reference dose of a chemical mixture (mg/kg/d)
the probability of injury, disease, or death under specific
circumstances
maximum acceptable risk level (uhitless)
individual cancer risk from the chemical mixture
relative source contribution
Registry of Toxic Effects of Chemical Substances
Science Advisory Board, U.S. Environmental Protection Agency
sister chromatid exchange
cancer slope factor, usually the upper 95 percent confidence limit
on the linear term (q^ in the multistage model
slope factor of a chemical mixture
size of the exposed population (number of persons)
the screening level concentration for a single contaminant in a
given fish species (mg/kg)
screening values of a given noncarcinogenic contaminant in a
given species of fish (mg/kg)
2,3,7,8-tetrachlorodibenzo-p-dioxin
toxicity equivalency concentration
toxicity equivalency factor
capable of causing physical defects in the developing embryo or
fetus
dose or exposure below which a significant adverse effect is not
expected
Toxics Release Inventory
xxiii
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GLOSSARY AND ABBREVIATIONS
uncertainty factors
URL
USDA
USGS
weight of evidence
WHO
WOE
WWW or WEB
one of several, generally 10-fold factors, used in operationally
deriving the RfD from experimental data. They are intended to
account for (1) the variation in sensitivity among the members of
the human population (intraspecies variability); (2) the uncertainty
in extrapolating animal data to humans; (3) the uncertainty in
extrapolating from data obtained in a study that is of less-than-
lifetime exposure to chronic exposure toxicity; (4) the uncertainty
in using LOAEL data rather than NOAEL data; and (5) uncertainty
generated by data gaps.
Universal Resource Locator
U.S. Department of Agriculture
U.S. Geological Survey
for carcinogens, this is a classification assigned to a chemical by
EPA, based on the types of data available concerning carcino-
genicity. On a scale of A to E, the classifications reflect the extent
to which the available biomedical data support the hypothesis that
a substance causes cancer in humans.
World Health Organization
weight-of-evidence
World Wide Web
XXIV
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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 second edition
of this document by the Research Triangle Institute (RTI) and Tetra Tech, Inc.
(EPA Contract No. 68-C3-0374). Patricia Cunningham of RTI was the contractor's
project manager. 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 (listed below). These 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. The workgroup was not involved in reviewing the
second edition because the basic risk assessment procedure had already been
approved. This second edition was issued primarily to update new toxicological
data for several analytes to incorporate existing supplemental information into the
body of the document, and to reformat the previous edition.
FISH CONTAMINANT WORKGROUP AND OTHER REVIEWERS
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
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
xxv
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ACKNOWLEDGMENTS
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
Other EPA Office Staff
Eletha Brady-Roberts
John Cicmanec
Michael Dourson
Susan Velazquez
Chon Shoaf
Jerry Stober
Charles Kanetsky
Milton Clark
Philip Crocker
Other Federal Agency Staff
Michael Bolger
Gregory Crame
Gunnar Lauenstein
Thomas Siewicki
Janice Cox
State Agency Staff
Anna Fan
Gerald Pollock
Richard Green
EPA/Office
EPA/Office
EPA/Office
EPA/Office
EPA/Office
EPA/Office
EPA/Office
EPA/Office
EPA/Office
EPA/Office
EPA/Office
EPA/Office
EPA/Office
EPA/Office
EPA/Office
EPA/Office
of Water
of Water
of Water
of Water
of Research
of Research
of Research
of Research
of Research
of Research
of Pesticide
of Pesticide
of Pesticide
of Pesticide
of Pesticide
of Pesticide
and Development
and Development
and Development
and Development
and Development
and Development
Programs
Programs
Programs
Programs
Programs
Programs
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
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
California
California
Delaware
xxvi
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ACKNOWLEDGMENTS
Joseph Sekerke
Tom Long
Dierdre Murphy
Jack Schwartz
John Hesse
Pamela Shubat
Gale Carlson
Alan Stern
Robert Tucker
Tony Forti
Luanne Williams
Martin Schock
Kandiah Sivarajah
Robert Marino
Kim Blindauer
Alan Anthony
Peter Sherertz
Ram Tripathi
Denise Laflamme
JimAmrhein
Henry Anderson
Native American Tribes
Neil Kmaicik
Ann Watanabe
John Banks
Clemon Fay
Florida
Illinois
.Maryland
Massachusetts
Michigan
Minnesota
Missouri
New Jersey
New Jersey
New York
North Carolina
North Dakota
Pennsylvania
South Carolina
Utah
Virginia
Virginia
Virginia
Washington
Wisconsin
Wisconsin
Great Lakes Indian Fish and Wildlife Commission
Columbia River Inter-Tribal Fish Commission
Penobscot Nation
Penobscot Nation
xxvii
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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 (Cunningham et al.,1990,1994). 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 (Cunningham et al., 1990).
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. 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.
XXVIII
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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 second edition because the basic risk
assessment procedures had already been approved. This second edition was
issued to update toxicological information for several of the target analytes,
incorporate existing supplemental information into the body of the document, and
reformat the previous edition.
This second edition provides risk assessors and managers with the most current
toxicological information for the target analytes and deletes carbophenothion as
a target analyte. This second edition has been reorganized to provide users with
• Detailed information on risk assessment methods including new 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)
• Concise information on how the monthly consumption limits tables were
developed and step-by-step instructions on how thesetables can be modified
to reflect local site-specific conditions for specific populations of concern
(Section 3)
• Monthly consumption tables for adults and children for all 25 target analytes,
and separate tables for women of reproductive age for methylmercury and
polychlorinated biphenyls (PCBs), that provide fish contaminant concentra-
tions ranging from no consumption limits (<6 meals per year) to safe con-
sumption (>3Q meals/month) limits (Section 4)
• A toxicological profile summarizing relevent toxicity data for each target
analyte including new toxicological information for methylmercury, PCBs,
inorganic arsenic, tributyltin, and polyaromatic hydrocarbons (PAHs) (Section
5)
• A brief discussion of geographic information system (GIS) mapping tools for
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
XXIX
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EXECUTIVE SUMMARY
(or from fish consumption data provided inAppendix 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).
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.
These documents are published in binder form so that they can be easily revised
and updated as significant new information becomes available. 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.
xxx
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1. INTRODUCTION
SECTION 1
INTRODUCTION
1.1 OVERVIEW AND OBJECTIVES
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
eating, 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).
In addition to the risks borne by the general population due to consumption of
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 noncommercially
caught fish and shellfish as a major source of protein in their diets. Within these
"at-risk" populations, pregnant women and children may be at even greater risk of
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1. INTRODUCTION
incurring adverse effects than other members of the populations, due to their
proportionally higher consumption rates and/or increased susceptibility to adverse
toxicological effects.
Fish contaminants vary widely in chemical structure and toxic properties. Potential
adverse health effects include cancer, chronic systemic effects, and develop-
mental and reproductive effects to name a few. The severity of these effects varies
with the exposure level and characteristics of the individual and may range from
relatively mild disease states to premature death. Recently attention has focused
on the developmental effects of chemical contaminants because studies con-
ducted over the past two decades have identified many environmental pollutants
as causing developmental abnormalities and other adverse reproductive out-
comes. For example, in developing a protocol for uniform sport fish consumption
advisories across the Great Lakes, the Great Lakes Sport Fish Advisory Task
Force used developmental and reproductive toxicity endpoints in calculating their
recommended consumption limits.
State, local, and Federal agencies and Tribal organizations currently use a range
of methods to estimate risks to human health from consumption of chemically
contaminated fish. Results of a 1989 survey of such methods (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 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 and revised in 1995),
Volume 2: Risk Assessment and Fish Consumption Limits (first released in 1994),
Volume 3: Risk Management (released in 1996), and Volume 4: Risk Communi-
cation (released in 1995). EPA recommends that the four volumes of this guidance
series be used together, since no one volume provides all the necessary informa-
tion to make decisions regarding the issuance of fish consumption advisories.
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1. INTRODUCTION
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. The target analytes listed in Table 1-1
were selected by EPA's Office of Water as particularly significant fish con-
taminants, 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.
The criteria for their selection are discussed in Volume 1, Section 4, of this series.
This second edition of Volume 2 makes refinements to the first edition including
major organizational changes in the presentation of the discussion of risk assess-
ment procedures and other information—inclusion of new information formerly
found in supplemental documents, the addition of revised toxicological summary
data on methylmercury and PCBs, and new toxicological data on arsenic, poly-
aromatic hydrocarbons (PAHs), and tributyltin.
1.2 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 consumers.
The last two sections provide summary information on the toxicology 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 of fish and shellfish, summarizes additions and revisions to this second edi-
tion, 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. New information has been incorporated into
Section 2 on population exposure assessment, including information on fish
consumption surveys and consumption patterns of various fisher groups, and dose
modification due to food preparation and cooking procedures. Additional informa-
tion on risk characterization has also been added.
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 calculation of risk-based consumption limits for carcino-
genic and noncarcinogenic effects using the appropriate cancer slope factor (q f)
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1. INTRODUCTION
Table 1-1. Target Analytes Recommended for Fish Sampling Programs
Metals
Arsenic (inorganic)
Cadmium
Mercury (methylmercury)
Selenium
Tributyitin
Organochlorine Pesticides
Chlordane, total (cis- 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 epoxideb
Hexachlorobenzene
Lindane (v-hexachlorocyclohexane; v-HCH)°
Mirexd
Toxaphene
Organophosphate Pesticides
Chlorpyrifos
Diazinon
Disulfoton
Ethion
Terbufos
Chlorophenoxy Herbicides
Oxyfluorfen
PAHs6
PCBs
Total Aroclorsf
Dioxins/furans9>h
PAHs = Polycyclic aromatic hydrocarbons.
PCBs = Polychlorinated biphenyls.
8 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.
0 Also known as v-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.
6 It is recommended that tissue samples be analyzed for benzo[a]pyrene, benz[a]anthracene,
benzo[6]fluoranthene, benzo[/c]fluoranthene, chrysene, dibenz[a,/?]anthracene, and indeno/X^S-ccflpyrene, and
that the order-of-magnitude relative potencies given for these PAHs in the EPA provisional guidance for
quantitative risk assessment of PAHs (U.S. EPA, 1993d) be used to calculate a potency equivalency
concentration (PEC) for each sample (see Section 5 of Volume 1). At this time, EPA's recommendation for risk
assessment of PAHs (U.S. EPA, 1993d) is considered provisional because quantitative risk assessment data
are not available for all PAHs. This approach is under Agency review and will be evaluated as new health
effects benchmark values are developed. Therefore, the method provided in this guidance document is subject
to change pending results of the Agency's reevaluation.
1 Analysis of total PCBs, as the sum of Aroclor equivalents, is recommended because of the lack of adequate
toxicologic data to develop consumption limits for individual PCB congeners (see Section 5). However, because
of the wide range of toxicities among different PCB congeners and the effects of metabolism and degradation
on Aroclor composition in the environment, congener analysis is deemed to be a more scientifically sound and
accurate method for determining total PCB concentrations.
9 Note: The EPA Office of Research and Development is currently reassessing the human health effects of
dioxins/furans.
h It is recommended that the 2,3,7,8-substituted tetra- through octa-chlorinated dibenzo-p-dioxins (PCDDs) and
dibenzofurans (PCDFs) be determined and a toxicity-weighted total concentration calculated for each sample
(Barnes and Bellin, 1989; U.S. EPA, 1987b) (see Section 5 of Volume 1). If resources are limited, 2,3,7,8-TCDD
and 2,3,7,8-TCDF should be determined at a minimum.
1-4
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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. Risked-based
Consumption
Limit Tables
5. Toxicological Profile
Summaries for
Target Analytes
6. Mapping Tools for
Risk Assessment and
Risk Management
Figure 1-1. Series Summary: Guidance for Assessing Chemical
Contamination Data for Use in Fish Advisories.
and reference dose (RfD). Information is also presented on calculation of con-
sumption limits for developmental effects in women of reproductive age and young
children. 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 con-
taminated with a single or multiple contaminants.
The monthly consumption limits for each of the 25 target analytes are provided in
Section 4. The tables list a number of alternative consumption limits for each target
1-5
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1. INTRODUCTION
analyte, based on different meal sizes, fish tissue contaminant levels, risk levels,
and toxicity endpoints. Specific consumption limits have been developed, and are
presented separately, for young children and adults in the general population. In
addition, consumption limits specifically targeted to women of reproductive age
have beep developed for methylmercury and PCBs based on new toxicological
information. Consumption limit tables have been added for arsenic, PAHs, and
tributyltin.
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, developmental toxicity, mutagenicity, carcino-
genicity, 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 (cancer slope factors and chronic
reference doses) 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. The toxicological profile summaries have been
modified for methylmercury and PCBs and new profile summaries have been
added for arsenic, PAHs, and tributyltin.
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 of fish 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 shown how 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 due to 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).
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1. INTRODUCTION
Risk assessment and risk management of contaminated fish are complex
processes due to 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).
1.3 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 the U.S. Food and Drug Administration (FDA)
were reviewed. Section 7 contains a complete listing of literature sources cited in
this document. Additional sources of information on risk assessment methods and
issues specifically related to fish risk assessment may be obtained from the
documents listed in Appendix A. These documents and scientific papers cover a
range of topics, from general risk assessment methods, to chemical-specific
toxicological data, to identification of chemical contaminant pathways.
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 the acknowledgments at the beginning
of this document. 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 basic risk assessment procedures
had already been approved. This second edition was issued primarily to update
new toxicological information for several analytes, to incorporate existing
supplemental information into the body of the document, and to reformat the
previous edition.
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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,c; IRIS, 1997).
EPA methods follow the outline developed in the National Academy of Sciences
(NAS) report entitled Risk Assessment in the Federal Government: Managing the
Process (NAS, 1983; see Figure 2-1). According to the NAS,
. . . 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. (NAS, 1983)
Readers may wish to consult the new NAS document, Science and Judgement in
Risk Assessment, which updates and expands the 1983 work (NAS, 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, orfoodborne) (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
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2. RISK ASSESSMENT METHODS
RESEARCH
RISK ASSESSMENT
RISK
MANAGEMENT
Laboratory and field
observations of adverse
health effects and
exposures to particular
agents
Hazard identification
(Does the agent cause
the adverse effect?)
Information on
extrapolation methods
for high to low dose
and animal to human
Dose-Response Assessment
(What is the relationship
between dose and incidence
in humans?)
Field measurements,
estimated exposures,
characterization of
populations
Exposure Assessment
(What exposures are
currently experienced
or anticipated under
different conditions?)
Development of
regulatory options
Evaluation of public
health, economic,
social, political
consequences of
regulatory options
Risk Characterization
(What is the estimated
incidence of the adverse
effect in a given
population?)
Agency decisions
and actions
Figure 2-1. Elements of Risk Assessment and Risk Management
(NAS, 1983).
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 identify the fish contamination target analytes of concern,
as described in Volume 1: Fish Sampling and Analysis (U.S. EPA, 1993a, 1995)
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 anaiytes 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
individuals. Population exposure assessments consider the distributions of
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2. RISK ASSESSMENT METHODS
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 been long 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, which address the following areas:
• Exposure assessment (U.S. EPA, 1992a)
• Carcinogenicity risk assessment (U.S. EPA, 1986a, 1996d)
• Mutagenicity risk assessment (U.S. EPA, 1986c)
• Developmental toxicity risk assessment (U.S. EPA, 1991 a)
• Assessment of female and male reproductive risk (U.S. EPA, 1996c)
• Health risk assessment of chemical mixtures (U.S. EPA, 1986d)
• 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 recent papers that discuss
risk assessment within EPA (e.g., Abernathy and 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 and Appendix A.
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2. RISK ASSESSMENT METHODS
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
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2. RISK ASSESSMENT METHODS
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 horn 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
(RAG). Readers may also wish to consult the open literature (e.g., Callahan et al.,
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 due to the extensive review conducted on the risk
values contained in the database. EPA's Health Effects Assessment Summary
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2. RISK ASSESSMENT METHODS
Tables (HEAST) are also frequently used (HEAST, 1992). 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.* ATSDR's Toxicity Profiles also provide detailed
toxicity data summaries.
2.2.1.2 Contaminant Data—
Information on the prevalence and measured concentrations of fish 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 in Fish (U.S. EPA, 1992b). In addition, substantial guidance
is provided in Volume 1 of this series on planning a sampling strategy and con-
ducting fish contaminant analyses (U.S. EPA, 1995).
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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2. RISK ASSESSMENT METHODS
• 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 II 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 of fish 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:
The RSC concept could be used in fish advisory activities. The
amount of exposure from fish consumed is determined along with
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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. See Appendix A for a listing of sources
of additional information.
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
from studies in experimental animals. There are often insufficient data in the
toxicological literature to fully characterize the toxicity of a chemical. Some types
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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., q^s and 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.
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 and Appendix A.
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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 D; 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 recent
NAS report, Pesticides in the Diets of Infants and Children (NAS, 1993), concluded
that 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:
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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 (NAS, 1993).
Although the focus of the NAS 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
NAS 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 cancer slope factors (q-|*s), reference doses, 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 1997. 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, 1992).
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
lexicologists, 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.
2.3.1 Carcinogenic Effects
EPA has recently proposed new guidelines for cancer risk assessment (U.S. EPA,
1996d). These guidelines have not been finalized yet but would supersede the
existing cancer guidelines (U.S. EPA, 1986d). 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
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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 (q^) 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, 1997). 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, 1996d).
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, 1996d) 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
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
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literature on this subject is not conclusive and readers may wish to review the NAS
report to obtain additional information (NAS, 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. Prenatal acute exposures are discussed in the DDT toxicological profile
summary in Section 5. Additional guidance on acute exposure risk assessment
methods may be provided in future revisions to this document. EPA welcomes
comments and recommendations on this and other methodologies.
Most toxicological 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 the less-toxic alternatives among a group of chemical options. LD5(^s 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
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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 the EPA's Office of Water). The Toxicological
Profiles developed by ATSDR contain Minimal Risk Levels (MRLs) for acute
effects for some contaminants. 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, 1997).
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, 1997). 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
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
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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, 1997). (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, 1997). 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 (as with
exposure limit estimating for developmental toxicity).
1. Identify the Most Appropriate NOAEL or LOAEL
The hierarchy for selection of a study described for developmental toxicity (Section
2.3.2.3) is also appropriate for use in identifying appropriate chronic toxicity
studies. 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 (IRIS, 1997).
2. Apply Relevant Uncertainty and Modifying Factors
The calculations for chronic systemic toxicity use the modifying and uncertainty
factors listed for developmental toxicity (see Table 2-1). In addition, an uncertainty
factor may be used when a chronic study is not available and a subchronic (e.g.,
90-day) study is used. This is generally a tenfold factor (Abernathy and Roberts,
1994; IRIS, 1997). 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,
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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 is relatively new (Abernathy and
Roberts, 1994). It has been developed because the dose-response data often
address a limited number of effects and may not adequately address effects of
major concern. 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,
1991 a); 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:
Table 2-1. Uncertainty Factors and Modifying Factors for Estimating Exposure
Limits for Developmental Effects
Uncertainty or Modifying Factor
General Comments
Standard
Value
Uncertainty factor: human (intraspecies)
"Uncertainty factor: animal to human
(interspecies)
Uncertainty factor: data gaps
Uncertainty factor: LOAEL to NOAEL
Modifying factor
Used to account for the variability of 10
response in human populations.
Used to account for differences in 10
responses between animal study
species and humans.
Used to account for the inability of 3 to 10
any study to consider all toxic
ehdpoints. The intermediate factor of
3 (1/2 log unit) is often used when
there is a single data gap exclusive of
chronic data (see IRIS, 1997).
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 (1997); and Jarabek et al. (1993).
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... 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 etal.,1992; U.S. EPA, 1989c).
The same type of concern regarding the completeness of a database is reflected
throughout the ATSDR Toxicological Profiles. For example, the profiles do not
provide an MRL for chemicals that have NOAELs and only LOAELs resulting in
severe effects.
The uncertainty and modifying factors are divided into the NOAEL or LOAEL to
obtain an estimated dose using the following equation:
RfD =
NOAEL or LOAEL
UF • MF
(2-1)
where
RfD
NOAEL or LOAEL
UF
MF
RfD or exposure limit for the target analyte
NOAEL from the selected study
multiplicative product of uncertainty factors
modifying factor.
This value is analogous to EPA's RfD. If an alternative exposure limit is calculated,
the results, in milligrams per kilogram per day, can be used in Equations 3.3 and
3.2, which are discussed in Section 3, to calculate fish meal consumption limits.
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
ATSDR Toxicological Profiles, IRIS, HDSB, and standard toxicological texts (all
are cited in the individual chemical discussions). The list of study characteristics
provided in Section 2.3.2.3 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.
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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
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 due to 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 20 percent
of the developmental defects observed in children are due to genetic causes, 10
percent to known factors, and 70 percent to 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, 1986e) and has been refined in the current Guidelines
for Developmental Toxicity Risk Assessment (U.S. EPA, 1991 a). 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, 1991 a). 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 for further information on developmental
toxicity risk and risk assessment.
EPA is working to incorporate new data on developmental and other types of
toxicity into the RfDs. The reader can use the information provided on individual
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target analytes (in Section 5) and the methodology discussed in this section to
calculate exposure limits based on their evaluation of the toxicological literature.
Section 5 identifies specific developmental outcomes and the associated dose-
response data (i.e., LOAELs or NOAELs) that can be used to carry out the
calculations. The reader may also wish to conduct a data search to identify any
recent data on the chemicals of major interest in their areas.
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 a). 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 1992e)
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 a).
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)
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
than only malformations and death, which were the primary effects considered
in the past.
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2. RISK ASSESSMENT METHODS
• 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 a).
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 a).
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 a). 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 a).
Definitions
There is no one consistent definition of developmental toxicity (U.S. EPA, 1986e).
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, 1986b) 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 ) structural abnormality, (3) altered
growth (defined below), and (4) functional deficiency.
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2. RISK ASSESSMENT METHODS
• 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, 1986b, 1991 a).
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 effect 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, 1986e) and, similarly, mutagenic effects can
be evaluated using criteria discussed in Guidelines for Mutagenicity Risk
Assessment (U.S. EPA, 1986c), as described in Appendix B.
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2. RISK ASSESSMENT METHODS
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 for the Health Assessment of Suspect Developmental Toxicants (U.S.
EPA, 1991 a) 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 pro-
spective studies have used the monitoring of pregnancy markers, such as human
chorionic gonadotropic (HCG) hormone, to identify very early post-implantation
pregnancy losses (see EPA, 1991 a 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, 1986e). 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 birth1 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, 1986e). The EPA health assessment guidelines describe
appropriate endpoints of maternal toxicity. One reason that identification of
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
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2. RISK ASSESSMENT METHODS
of greatest concern because these dynamics suggest that developing individuals
are more sensitive or selectively affected (U.S. EPA, 1986e). 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, 1986e).
Methods for Estimating Exposure Limits
This section was not designed to provide detailed guidance on conducting dose-
. response evaluations. Rather, it provides a more detailed discussion of the EPA
method to calculate RfDs, which is presented in Section 2.3.2.2. This method can
be used by the reader to estimate exposure limits for developmental effects as
necessary. As previously discussed, the major steps are identification of the most
appropriate NOAEL or LOAEL and application of relevant uncertainty factors and
modifying factors.* This discussion assumes a familiarity with basic concepts in
epidemiology, toxicology, and human biology. Guidance is also provided in the
discussions of individual target analytes (Section 5.3) on selection of a sensitive
health endpoint or study and use of uncertainty and modifying factors.
1. Identify Most Appropriate NOAEL or LOAEL
The approach discussed in this section uses NOAELs and LOAELs in a manner
analogous to that used for the development of chronic toxicity RfDs. The EPA
guidance on developmental toxicity (U.S. EPA, 1991 a) also discusses the use of
a benchmark dose to evaluate toxicity. This approach employs a different method
of evaluation than that previously described under chronic exposure in Section
2.3.2.2. The benchmark approach uses the response rate as a critical factor (e.g.,
the dose effective in 10 percent of the study subjects). Such an approach requires
more extensive information than is available for most target analytes. It is
recommended that the reader review the 1991 guidance on developmental toxicity
risk assessment, which provides extensive specific guidance on the evaluation and
selection of various types of developmental toxicity studies (U.S. EPA, 1991 a). The
1991 guidelines provide a scheme for categorization of health-related data, which
includes descriptions of sufficient evidence and insufficient evidence for dose-
response evaluations. The guidelines recommend that a dose-response evaluation
not be conducted unless there is sufficient evidence. To evaluate developmental
toxicity, data from human studies may be used. However, for most chemicals,
human study data are not available and toxicity studies in animals are used (U.S.
EPA, 1987a, 1991 a). EPA's Office of Health Effects Assessment (OHEA) may also
be consulted for guidance on obtaining additional information and identifying
existing databases on developmental toxicants.
Characterization of the database is also an important step. However, it is assumed in this document that
an abbreviated approach will be taken to estimating exposure limits. If the summary data provided in this
work or taken from other sources are used, it will not be possible to fully characterize and categorize the
database. (See U.S. EPA, 1991 a, pp. 63816-63817.)
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2. RISK ASSESSMENT METHODS
Exposure limits may be estimated using the NOAEL or LOAEL obtained from
toxicological studies of animals and humans or epidemiological studies of humans.
The NOAEL, usually expressed in mg dose per kg body weight of the subject per
day is the highest dosage given to the animals over their lifetime that results in no
observable adverse effects. When a NOAEL is not available, the lowest dose at
which an adverse effect was observed is used. Often there are several NOAELS
and LOAELs for a chemical; selection of the most appropriate value is a judgment
based on the quality of the studies, sensitivity of the health endpoint and test
species, and numerous other factors. 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.
• 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.
It is necessary to consider overall study quality and study design in selecting the
most appropriate study or studies. The reader should refer to the 1991 Guidelines
for Developmental Toxicity Risk Assessment for further details (U.S. EPA, 1991a).
2. Apply Relevant Uncertainty and Modifying Factors
Once a LOAEL or NOAEL is selected, the value obtained (in mg/kg/d) is divided
by factors to account for the various types of uncertainty inherent in estimating a
threshold for developmental effects. These factors, referred to by EPA as
uncertainty factors and modifying factors, are summarized in Table 2-1, which was
adapted from a discussion of RfD development in Abernathy and Roberts (1994).
Many developmental toxicity studies use a brief dosing period during gestation
(although use of a study with a single dose is not recommended). An uncertainty
factor is usually not added for the short duration of the study under these cir-
cumstances. This differs from the calculation of exposure limits based on chronic
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2. RISK ASSESSMENT METHODS
exposure toxicity (discussed previously in this section) when an uncertainty factor
is typically applied for the use of a less-than-lifetime study.
The total product of all of the uncertainty and modifying factors may range widely
depending on the types of studies available. If a chronic human epidemiologic
study is available, the uncertainty factor may be as small as 1. However,
uncertainty factors of 10,000 have been used (IRIS, 1997). While the uncertainty
factors address specific concerns, the modifying factors cover 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 a sensitive endpoint. The default value for a modifying
factor is 1. The uncertainty factor that deals with data gaps is relatively new
(Abernathy and Roberts, 1994). It has been developed because the dose-
response data often address a limited number of effects and do not adequately
address effects of major concern. In some cases there are a number of studies,
but the focus of analysis is narrow and insufficiently sensitive. In other cases, there
is not a sufficient number or breadth of studies (see Dourson et al., 1992, for
experimental support of this database factor). Other reasons for applying a
modifying factor are discussed in the specific developmental toxicity guidance
(U.S. EPA, 1991 a); these include data on pharmacokinetics or other considera-
tions that may alter the level of confidence in the data.
The uncertainty and modifying factors are divided into the NOAEL or LOAEL to
obtain an RfD using Equation 2-1 (Section 2.3.2.2). If an exposure limit is cal-
culated for developmental toxicity, the results, in mg/kg/d, can be used in Equa-
tions 3-3 and 3-2 discussed in Section 3, to calculate fish consumption limits.
Examples of how this is carried out are provided in Section 3.
As discussed above, it is necessary to have a full characterization of the
uncertainties and assumptions incorporated in fish consumption limits. Assump-
tions and uncertainties associated with dose-response assessment are discussed
in Sections 2.3.5 and 5.1.1.12. As a point of reference, EPA has estimated that the
RfDs that they develop have an uncertainty spanning approximately 1 order of
magnitude (U.S. EPA, 1987a). A description of the variability in dose-response
results and their impact on fish consumption limits and descriptions of the data
gaps, study limitations, and assumptions are also important in providing a context
for fish consumption limits based on developmental 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 Section 2.3.5 to identify
major sources of uncertainty. In addition, the list of study characteristics provided
previously in this section may be useful for identifying data gaps and sources of
uncertainty. If this document is the only source consulted for dose-response data,
it should be noted that the literature review conducted for the development of
values was limited to secondary sources such as ATSDR Toxicologicai Profiles,
IRIS, HDSB, and standard toxicological texts (all are cited in the individual
toxicological profile summaries). The inclusion of this type of information in the risk
management process following risk assessment will provide a better overall
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2. RISK ASSESSMENT METHODS
understanding of the limitations and uncertainties inherent in the fish consumption
limits.
The 1991 developmental toxicity risk assessment guidelines provide a scheme for
categorization of health-related data, including descriptions of sufficient and
insufficient evidence for dose-response evaluations. The guidelines recommend
that a dose-response evaluation not be conducted unless there is sufficient
evidence (U.S. EPA, 1987a, 1991 a). The reader is referred to this source for
additional information on all aspects of risk assessment for developmental toxicity.
EXAMPLE
The chemical group polychlorinated biphenyls (PCBs) was chosen as an
example of how an estimated exposure limit for developmental effects can be
developed for target analytes.
It is advisable to conduct a thorough literature search to identify all relevant
studies. The summaries of dose-response and other toxicity data provided in
Section 5 provide an overview; however, it is advisable to seek additional data,
including any newly released information, whenever practical. An abbreviated
approach to estimating an exposure limit, using the information provided in this
guide, is discussed below.
In addition to the data in Section 5 specifically discussing developmental
toxicity, it is useful to review other relevant data. This includes chronic toxicity
and carcinogenicity, including especially reproductive system toxicity and other
organ toxicities that are similar to, or affect the same system as, that observed
in developmental toxicity studies. All other sections of the target analyte
toxicological profile summary may also have a bearing on understanding and
interpreting the results of developmental toxicity studies. They may support or
refute the results observed or point out potential data gaps (e.g., organ toxicities
observed in numerous studies of adult animals but not evaluated in develop-
mental toxicity studies).
It is especially necessary to review any discussions of reproductive system
toxicity in the Chronic Toxicity section of a target analyte discussion. This may
have a bearing on the interpretation of developmental toxicity study results. For
example, alteration in hormonal balances, structural changes in the reproductive
system, and other adverse effects may modify the ability to maintain pregnancy.
This could lead to a reduction in the number or size of litters or other impact on
pregnancy outcome. These factors would need to be considered when review-
ing developmental toxicity studies that identify effects such as increased fetal
resorptions, fetal death, reduced litter size, and related effects because these
effects could arise from damage to the mother rather than the offspring.
(continued'
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2. RISK ASSESSMENT METHODS
EXAMPLE (continued)
There are numerous other effects on reproductive toxicity that may affect
interpretation of developmental toxicity study results. The reader may wish to
consult texts on this subject for further information. See the relevant repro-
ductive system toxicity discussion for PCBs in Section 5.7.
It may also be helpful to survey the available information on related chemicals
(e.g., structural relatives of PCB would include organochlorine pesticides). This
may provide general information on effects that are common to several or all
members of a chemical group. Such findings lend support to conclusions
regarding toxicity. In addition, studies on related chemicals may have explored
effects anticipated (based on adult studies in the chemical of concern) in
developing individuals but not evaluated in developmental studies on the
chemical of concern. This provides useful information for qualitatively evaluating
potential toxicity and may point out critical data gaps.
For Aroclor 1016, the RfD is based on developmental effects seen in monkeys.
A NOAEL of 0.001 mg/kg/d was established and an uncertainty factor of 100
was applied (3 for sensitive individuals, 3 for interspecies extrapolation, 3 for
database limitations, and 3 for extrapolation from subchronic to chronic). This
results in an RfD of 7 x 10"5 mg/kg/d
Estimated
Exposure =
Limit 100
nn/v_
0'°°7
= 7 x MT5 mg/kg/d
(2-2)
where
0.007 = NOAEL from the selected study
100 = uncertainty factor
A discussion of uncertainty, assumptions, and data gaps should be a part of
information supporting an estimated exposure limit. This information can include
a summary of the various sources of uncertainty described in Sections 2.3.5 and
5 of this document, information included in the target analyte discussion, and
any other information the reader feels would be useful in characterizing
uncertainty.
(continued)
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2. RISK ASSESSMENT METHODS
EXAMPLE (continued)
The discussion of data gaps in Section 5.7 includes a list of the types of studies
needed based on an interagency review of the available data. Most major
categories of uncertainty are covered by this list of studies. The reader may wish
to elaborate on why certain studies are needed (e.g., pharmacokinetic studies
to generate better information on bioaccumulation, body burden, accumulation
in breast milk).
Sources of Additional Information on Developmental Toxicity
The primary source the reader is referred to for additional information on con-
ducting risk assessment for developmental toxicity is: Guidelines for Develop-
mental Toxicity Risk Assessment (U.S. EPA, 1991 a). In addition, there are 165
citations listed in the Guidelines that cover a broad spectrum of literature on
developmental toxicity and risk assessment. The reader may wish to consult these
sources for additional guidance on this topic.
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
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
B.
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
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2. RISK ASSESSMENT METHODS
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, 1986d). The
guidelines advise the use of the additive approach when data are available only
on individual mixture components. 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. Similarities in the toxicity characteristics of
organochlorinated pesticides and organophosphate pesticides are discussed in
Appendix C. It does not address chemicals with dissimilar actions.
The 1986 Guidelines also address circumstances when data are available on
antagonistic or synergistic interactions. They state that "information must be
assessed in terms of both its relevance to subchronic or chronic hazard and its
suitability for quantitatively altering the risk assessment." These two criteria are
essential for selection of interactive data applicable to quantitative risk
assessment. However, the criteria preclude the use of most interactive data in risk
assessments of long-term exposures because many interactive studies focus on
short-term exposure. An additional complication is introduced to this type of
analysis for mixtures containing more than two chemicals. For those groups, it is
necessary to ascertain whether the presence of additional chemicals in the mixture
will alter any known interactions between any two chemicals having interactive
data (U.S. EPA, 1986d).
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, 1986d). 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. Much of the interactive information
included in the toxicological profiles in Section 5 was obtained from the ATSDR
Toxicological Profiles for various chemicals. Readers are encouraged to consult
these ATSDR documents if they require interactive data.
The information obtained from ATSDR and/or 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 microsomai 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.
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2. RISK ASSESSMENT METHODS
EPA has recently 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 oflhe 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 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 fully evaluate 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., for carcinogenicity, 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 [EPA, 1986a,c,e; 1996d]).
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
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2. RISK ASSESSMENT METHODS
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 a-2-globulin may not be applicable to humans).
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 [nongeneticj
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, 1997). 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
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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,
1997).
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
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's new
cancer guidelines (which have not been finalized as of this writing) propose using
straight line extrapolation (U.S. EPA, 1996d).
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 certainly 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).
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2. RISK ASSESSMENT METHODS
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
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 A.
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.
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2. RISK ASSESSMENT METHODS
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
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, 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, 1995),
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) pro-
cedures, 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 are also included.
Information on contaminant distributions in different types of fish and fish tissues
and across geographical 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
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2. RISK ASSESSMENT METHODS
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 of fish contamination that, combined
with demographic information, help target exposed fisher populations for additional
risk communication and 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 contamination of fish 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:
Cm.CR
BW
(2-3)
where
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
(Cm) and on human consumption patterns (CR/BW) to estimate exposure (Em) for
hypothetical individuals within given populations (see Equation 2-1). Conversely,
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2. RISK ASSESSMENT METHODS
the consumption limits described in Section 3 and provided in Section 4 use the
data on known chemical residues in fish (Cm) combined with dose-response data
(q.,*s 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).
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 (C j 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 noncommercially
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 2-3 uses three parameters to calculate individual exposure (E J to fish
contaminants from noncommercially caught fish: consumption rate (CR), consumer
body weight (BW), and contaminant concentration (C J. Equations 3-1, 3-2, and
3-3 in Section 3 also use body weight and contami.nant concentration and meal
size (MS) in developing consumption limits. With the exception of C^ which is
determined through sampling and analysis programs, these parameters are
discussed below.
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2. RISK ASSESSMENT METHODS
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
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 (NHANES II).
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 repro-
ductive 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
age 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.
Table 2-2. Mean Body Weights of Children and Adults
Mean Body Weight (kg)
Age Group
Males
Females
Males and Females
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
Readers are encouraged to use values that average together male and
nonpregnant female body weights when assessing exposure to the general adult
population. Where consumption rates are known to differ significantly between
men and women, however, readers may wish to make gender-specific exposure
assessments and use unaveraged gender-specific body weight estimates. When
certain developmental toxicants are of concern, readers are encouraged to make
separate exposure assessments for children and women of reproductive age.
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
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 cooked fillets. Still others do
not specify whether the value is based on cooked 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,
1995). 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 (1988c). A discussion of fish consumption surveys is
provided in Appendix D.
EPA has developed meal size estimates for both adults and children under 4. The
general adult population includes all adult men and women. Children eat smaller
portions than adults, although they may consume significantly more fish per unit
body weight. Women of reproductive age were assumed to eat proportionally (per
kg body weight) the same amount of fish per meal as other adults.
EPA suggests using a default value of 8 oz (227 g) of cooked fish fillet per 70-kg
consumer body weight as an average meal size for both the general adult
population and for women of reproductive age 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.
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2. RISK ASSESSMENT METHODS
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.
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. Consequently, fish consumption patterns and
methods for evaluating the resulting risks are presented in a new version of the
Guidelines for Exposure Assessment, which EPA is currently finalizing. 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.
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2. RISK ASSESSMENT METHODS
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, 1986d) and EPA's proposed cancer guidelines do not
address this issue (U.S. EPA, 1996d). 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 the acceptable risk listed in each column (i.e., 1 in
10,000,100,000, and 1 million).
The likelihood of occurrence of noncarcinogenic effects associated with chronic
exposure is evaluated through the use of RfDs (as discussed in Section 2.3).
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. It is anticipated that subsequent
revisions of this document will more fully characterize acute exposure (see Section
2.3 for a brief discussion).
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
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2. RISK ASSESSMENT METHODS
developmental effects including stillbirths have been associated with exposures
to high levels of pesticides in foods. Information is provided in a recent NAS report
on developmental toxicity on special characteristics of infants and children that
cause their exposures and risks to differ from those of adults (NAS, 1993). If very
high exposures are likely to occur, State agency staff are encouraged to consider
this exposure scenario in more detail.
Risk managers may wish to use other averaging periods (e.g., 7 days, 10 days, or
14 days) for developing short-term consumption limits to better represent
vacationers involved in recreational fishing. Using a 10-day averaging period for
short-term exposures can be justified for several reasons. Ten days is one of the
averaging periods used by the EPA Office of Water in developing Health
Advisories for drinking water. It is also relevant to the short time period often
considered critical for exposure to developmental toxicants (i.e., there may be a
brief window of time during which adverse effects can be induced by toxicants).
This time period also corresponds to a typical vacation period. Although some fish
consumption advisories use 3 weeks as an exposure period to describe
recreational fish consumption (Minnesota Department of Health, 1992; Missouri
Department of Health, 1992), no evidence was found to support it as a more
accurate period than 10 days. Note: Vacationers may identify better with 2-week
periods than with 10-day periods (Shubat, 1993a). For this reason, readers
intending to develop advisories based on 10-day fish consumption may want to
consider expressing consumption limits in terms of a 2-week vacation period.
instead. As an example, a 10-day meal consumption limit table has been
developed for chronic systemic health effects of chlordane is included in Section
3 (Table 3-13). Procedures used to calculate 10-day or other short-term time-
averaged consumption limits are described in Section 3.3.6 for those risk
managers who want to use these shorter time-averaging periods to better
communicate appropriate consumption limits to recreational fishers. Note: Those
recreational fishers who catch and freeze large quantities of fish to eat later might
be considered seasonal or subsistence fishers, depending on the extent of their
catch and subsequent consumption.
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 2-3 can be modified, as
follows, to consider consumption of multiple species:
BW
(2-4)
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2. RISK ASSESSMENT METHODS
where
Em,j =
Cm,j =
PJ =
individual exposure to chemical contaminant m from ingesting fish
species j (mg/kg/d)
concentration of chemical contaminant m in the edible portion of fish
species j (mg/kg)
consumption rate of fish species j (kg/d)
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 (Honstead et al., 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 2-4) 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
if possible. Consumption patterns are discussed in greater detail in Appendix D.
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2. RISK ASSESSMENT METHODS
Table 2-3. Categories of Information Necessary for a Population Exposure
Assessment
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)
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 D. 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 E.
Category 7, which deals with multimedia exposure assessment, may be very
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
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2. RISK ASSESSMENT METHODS
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. Decision-makers 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 decision-makers. 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 decision-makers 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.
EPA is currently finalizing a new version of the Guidelines for Exposure
Assessment. Information regarding the new guidelines will be provided in future
editions of this series.
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.
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 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
Typical
Percentile
of
Exposure
90%] 95% 98% 99%
99.9%
High End of Exposure
Souroe: Hablcht, 1992.
Figure 2-2. Schematic of Exposure Categories in Upper Half of a Normal Population
Distribution.
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).
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 bioavaiiability, 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.
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2. RISK ASSESSMENT METHODS
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 separately
(Habicht, 1992). The subpopulatiqns may differ from the rest of the population by
virtue of their activities, age, sex, lifestyle, economic factors, residence, diet,
cultural patterns, physiology, or other factors (Habicht, 1992).
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.
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 ((J.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 (U.S. EPA, 1992a). If an upper limit
rather than a best estimate is used, express this clearly with the exposure
estimate.
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2. RISK ASSESSMENT METHODS
• 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.
• 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 conser-
vative 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 oyerestimation 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
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2. RISK ASSESSMENT METHODS
• 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
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 arid 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. The contaminant concentrations in fish
tissue below the LOD are shaded where appropriate in Section 4 tables to alert
risk assessors to uncertainty in measuring these low concentrations in fish tissue.
2.4.5.2 Dose Modifications Due to Food Preparation and Cooking—
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 may be found in Appendix E.
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. In addition, the
consumption limit tables take differences in body weight into account by scaling
meal size to body weight (e.g., 8-oz meal per 70-kg body weight).
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
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2. RISK ASSESSMENT METHODS
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
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.
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2. RISK ASSESSMENT METHODS
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
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 who may subsist on noncommercially caught fish year-
round, including Native Americans and certain recent immigrants accustomed to
self-sufficiency and fishing (particularly Asian-Americans) and economically
disadvantaged populations may be particularly 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
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r
2. RISK ASSESSMENT METHODS
from a variety of waterbodies. Single-species consumption limits also do not
address related species that may be contaminated but were not sampled. Such
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. Consequently, there is no acute exposure methodology
recommended (that would correspond to bolus doses; see Section 2.3) in this
document. In subsequent editions, this type of information may be included.
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 2-4 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, drinking water, inhalation, or dermal contact) to experience negative health
effects and/or increased cancer incidence, even if they1 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 an exceedence of 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 of fish 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 the total allowable daily dose (e.g., 10, 20, or 30 percent). This allocates to the
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2. RISK ASSESSMENT METHODS
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:
ET =
cm-CR
BW
Ew +. EF
(2-5)
where
CR
BW
Ew
E
E =
exposure from all sources (mg/kg/d) to contaminant (m)
concentration in the edible portion of fish (mg/g)
mean daily consumption rate of fish (g/d)
average body weight of the group (kg)
exposure from air sources (mg/kg/d)
exposure from water sources (mg/kg/d)
exposure from nonfish food sources (mg/kg/d)
exposure from other sources (e.g., 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
contaminants, with a proportional loss of fillet weight, will not change the
concentration, which is expressed in milligrams of contaminant per kilogram of fish
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2. RISK ASSESSMENT METHODS
(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 3-8.
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 whom they are designing the fish advisories to protect, 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 of fish that are caught, the
quantity of fish consumed, and the method of preparation and cooking used. In
some cases, other factors willalso affect exposure (e.g., seasonal changes in
contaminant levels, the age of the fish). For purposes of risk assessment,
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2. RISK ASSESSMENT METHODS
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2. RISK ASSESSMENT METHODS
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 regarding 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 E 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 are to be developed
for this group, information regarding 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 (NAS, 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 their differences from the
larger 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.
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2. RISK ASSESSMENT METHODS
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
regarding 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 of fish
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)
• 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
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2. RISK ASSESSMENT METHODS
• 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 regarding 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. Decision-makers 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.
See Appendix F 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
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2. RISK ASSESSMENT METHODS
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
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). State agencies
should evaluate their approach to this issue based on their review of the available
literature.
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).
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2. RISK ASSESSMENT METHODS
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.
When different exposure levels occur, this calculation is repeated for each
exposure group. The total risk is the sum of the risks at each exposure level:
total risk = risk at exposure level a + risk at exposure level b + ...
risk at exposure level 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 calculated so that there is little probability of an adverse health
effect occurring due to chronic exposure to chemical concentrations below the
RfD. Exceedence of the RfD implies there may be some risk of the adverse health
effect occurring; however, the magnitude of risk and severity of the effect are not
quantified by this approach.
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
where
exposure = total exposure to a single contaminant from all sources (mg/kg/d)
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:
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2. RISK ASSESSMENT METHODS
noncarcinogenic risk = population with exposure greater than the RfD.
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 regarding 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 decision-makers to be in need of fish 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 with
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
consumption advisory. For noncarcinogenic effects, exposures greater than the
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2. RISK ASSESSMENT METHODS
RfD by a factor of 1,10, or some other values 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 regarding 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 regarding 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).
Due to the important nature 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
encourages States, 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 people
served.
Some contaminants have very well-supported risk values. Others have values that
are based on limited data, or the data suggest risks may occur that are not
quantitatively definable at this time. In these cases, State agencies may choose
to consider risks from a more health conservative viewpoint. Alternatively, the risk
values could be modified by State agencies to calculate risks reflecting these
concerns. The use of risk data, and its impact on populations and how fish
advisory programs are designed, is at the discretion of State, local, and Tribal
agencies. Some chemicals have information available, but the data are not
sufficient to quantify the risk to sensitive subpopulations. For example, DDT is
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2. RISK ASSESSMENT METHODS
thought to cause effects in two sensitive populations for which no quantitative
dose-response data exist. Based on animal studies, DDT may cause cardiac
sensitization leading to ventricular fibrillation and death (Hayes, 1982). However,
there are insufficient data to relate this quantitatively to human populations that
have cardiac disorders or other stresses that could make them especially
susceptible to these effects. DDT may also cause disturbances in the normal
reproductive system of females (Hayes, 1982), although, again, no quantitative
dose-response data exist for these effects. Another example is that certain
chemicals (such as organophosphate pesticides) are metabolized more slowly in
certain individuals with specific enzyme deficiencies than in the general population
(Hayes, 1982). The dose-response data for these effects are not available. In all
these cases, the probability or the magnitude of the response that may result from
a given dose cannot be quantified adequately. Information on these effects of DDT
on sensitive subpopulations and other chemicals that may result in risk to certain
sensitive populations are described in Section 5. Readers can use this information
in several ways. Sections 2.3.2.2 and 2.3.2.3 briefly discuss methods for applying
modifying factors to the toxicity reference values in cases where there are
insufficient data to quantify risks. Readers may also choose to consider these
effects qualitatively when developing consumption limits and/or making risk
assessments. In addition, consumption advisories can contain a discussion of
suspected effects of a given chemical for particular subpopulations, so that
individuals with such conditions can make informed decisions regarding their
consumption of fish. Finally, readers may wish to apply an additional safety factor
to the meal advice aimed at such populations to provide some measure of
additional protection.
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
wisti 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 ah
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
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2. RISK ASSESSMENT METHODS
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 provides methods to summarize population exposure arid 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.
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 regarding 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.
2-65
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2. RISK ASSESSMENT METHODS
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2. RISK ASSESSMENT METHODS
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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 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
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2. RISK ASSESSMENT METHODS
example, sport fishers and subsistence fishers may use the same waterbody but
have different risks based on their varied consumption habits.
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
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.
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2. RISK ASSESSMENT METHODS
Table 2-7 also provides data entry areas for three populations surrounding a
waterbody (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 decision-makers 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
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
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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 con-
sumption limits for a single contaminant in a multiple species diet or for multiple
contaminants causing the same chronic health effect endpoints are also dis-
cussed. Species-specific consumption limits are calculated in kilograms per day
and converted to allowable fish meals in ounces that may be consumed per
month. This approach is taken because consumers tend to think of fish consump-
tion in terms of meals rather than in terms of grams or ounces.
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
3-1
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3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
sample types). However, for populations that ingest whole fish, consumption
values corresponding to whole fish contaminant concentrations are more appro-
priate. Fish consumption patterns are discussed in more detail in Appendix D.
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, although
Appendix A provides a list of references and government agencies that may be of
assistance in quantifying these other sources of exposure.
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 noncarcinogenic
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 range of risk levels from 1 in 10,000 (10'4) to 1 in 1 million
(10"6). Equations 3-1 and 3-2 were used to calculate risk-based consumption limits
for the 14 target analytes with cancer slope factors (see Table 3-1), based on an
assumed 70-year exposure. A 70-year 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 Hfespan, 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. However, readers
may wish to calculate consumption limits for specific cohorts (e.g., children,
individuals exposed to other carcinogens) based on their interpretation of the
lexicological literature and local conditions.
3-2
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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)b
Cadmium
Mercury (methylmercury)0
Developmental11
Chronic systemicd
Selenium6
Tributyltin
Organochlorine Pesticides
Total chlordane (sum of cis- and frans-chlordane,
cis- and frans-nonachlor, and oxychlordane)'
Total DDT (sum of 4,4'- and 2,4'-
isomers of DDT, DDE, and ODD)9
Dicofolh
Dieldrin
Endosulfan (1 and ll)h
Endrin
Heptachlor epoxide
Hexachlorobenzene
Lindane (v-hexachlorocyclohexane; v-HCH)'
Mirex
ToxaphenehJ
Organophosphate Pesticides
Chlprpyrifos
Diazinonh
Disulfoton
Ethion
Terbufosh
Chlorophenoxy Herbicides
Oxyfluorfen
PAHsk
PCBs
Total PCBs (sum of Aroclors)
Developmental
Chronic systemic
Dioxins/furans1
NA = Not available in EPA's Integrated Risk
Information System (IRIS, 1997).
PAH = Polycyclic aromatic hydrocarbon.
Chronic RfDa q.*a
(mg/kg/d) (moL/dV1
3x10-4 1.5
1x10'3 NA
1x10'4 NA
1x10'4 NA
5x10"3 NA
3x10'5 NA
6x10"5 1.3
5x1Q-4 0.34
1.2 x1Q-3 0.34
5x10'5 16
6x10'3 NA
3x10'4 NA
1.3x10~5 9.1
8x10"4 1.6
3x10'4 1.3
2x10'4 1.8
3.6 x10'4 1.1
3x10'3 NA
9x10"5 NA
4x10'5 NA
5x10'4 NA
1.3 x10'4 NA
3x10'3 1.28x10'1
NA 7.3
2x1Q-5(1254)
2x10-5(1254) 2.0
NA 1 56 x 1 05
PCB = Polychlorinated biphenyl.
RfD = Oral reference dose (mg/kg/d).
q.,*SF = Cancer slope factor (mg/kg/d)"1.
(continued)
3-3
-------
Table 3-1 (continued)
* Unless otherwise noted, values listed are the most current oral RfDs and q^ in EPA's IRIS database (IRIS,
1997).
b Total inorganic arsenic should be determined.
0 Because most mercury in fish and shellfish tissue is present primarily as methylmercury (MAS, 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.
d Note: The EPA reevaluated the RfD for methylmercury in 1995, primarily because of concern about evidence
that the fetus is at increased risk of adverse neurological effects from exposure to methylmercury (Marsh et al.,
1987; Piotrowski and Inskip, 1981; MAS, 1991; WHO, 1976,1990). On May 1,1995, IRIS was updated to include
an oral RfD of 1x10'4 mg/kg/d based on developmental neurological effects in human infants. This oral RfD of
1x10"4 mg/kg/d is considered protective for chronic systemic effects of methylmercury among the general adult
population, women of reproductive age, and children. Additional toxicological information on methylmercury is
provided in Section 5.
6 The RfD for selenium is the IRIS (1997) value for selenious acid.
1 The RfD and q/ values listed are derived from studies using technical-grade chlordane (purity -95%) or a 90:10
mixture of chlordanerheptachlor or analytical-grade chlordane (IRIS, 1997). No RfD or SF values are given in
IRIS (1997) for the cis- and frans-chlordane isomers or the major chlordane metabolite, oxychlordane, or for the
chlordane impurities cis- and frans-nonachlor. It is recommended that the total concentration of cis- and trans-
chlordane, cis- and frans-nonachlor, and oxychlordane be determined.
0 The RfD value listed is for DDT. The SF value is for DDT or DDE; the SF value for ODD is 0.24. The U.S. EPA
Carcinogenicity Assessment Group recommended the use of SF = 0.34 for any combination of DDT, DDE, ODD,
and dicofol (Holder, 1986). It is recommended that the total concentration of the 2,4'- and 4,4'-isomers of DDT
and its metabolites, DDE and ODD, be determined.
h The RfD value listed is from the Office of Pesticide Program's Reference Dose Tracking Report (U.S. EPA,
1996b).
' IRIS (1997) has not provided a cancer slope factor for lindane. The q,* value listed for lindane was calculated
from the water quality criteria (0.063 ug/L) (U.S. EPA, 1992i).
1 The RfD value has been agreed upon by the Office of Pesticide Programs and the Office of Water.
k The SF value listed is for benzo[a]pyrene. Values for other PAHs are not currently available in IRIS (1997). It is
recommended that tissue samples be analyzed for benzo[a]pyrene, benz[a]anthracene, benzo[fc]fluoranthene,
benzoWfluoranthene, chrysene, dibenz[a,/7]anthracene, and indeno/X2,3-cc(]pyrene, and that the order-of-
magnitude relative potencies given for these PAHs in the EPA provisional guidance for quantitative risk
assessment of PAHs (U S. EPA, 1993d) be used to calculate a potency equivalency concentration (PEC) for
each sample (see Section 5.3.2.3 of Volume 1). At this time, EPA's recommendation for risk assessment of
PAHs (U S EPA 1993d) is considered provisional because quantitative risk assessment data are not available
for all PAHs This approach is under Agency review and will be evaluated as new health effects benchmark
values are developed. Therefore, the method provided in this guidance document is subject to change pending
results of the Agency's reevaluation.
' The SF value listed is for 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) (U.S. EPA, 1986e). It is recommended that
the 17 2 3 7 8-substituted tetra- through octa-chlorinated dibenzo-p-dioxins and dibenzofurans be determined
and a toxicity-weighted total concentration be calculated for each sample, using the revised interim method for
estimating Toxiclty Equivalency Concentrations (TECs) (Barnes and Bellin, 1989; U.S. EPA, I991c). If resources
are limited the 2,3,7,8-TCDD and 2,3,7,8-TCDF congeners should be determined at a minimum.
3-4
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3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
3.2.1.1 Calculation of Daily Consumption Limits —
Equation 3-1 calculates allowable daily consumption of contaminated fish based
on a contaminant's carcinogenicity, expressed in terms of kilograms of fish
consumed per day:
ARL • BW
(3-1)
where
lim
ARL
BW
CRllm = maximum allowable fish consumption rate (kg/d)
= maximum acceptable individual lifetime risk level (unitless)
= consumer body weight (kg)
c\f = cancer slope factor, usually the upper 95 percent confidence limit
on the linear term in the multistage model used by EPA [(mg/
k9-d) 1. (see Section 2 for a discussion of this value)
Cm = measured concentration of chemical contaminant m in a given
species of fish (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. Four meal sizes were used to develop the consumption limit tables for
carcinogens in Section 4: 0.114 kg (4 oz), 0.227 kg (8 oz), 0.341 kg (12 oz), and
0.454 kg (16 oz). Note: Although all calculations are in units of kilograms, meal
sizes are converted to ounces in the tables in Section 4 for ease of risk communi-
cation to U.S. consumers. The conversion rate from ounces to kilograms is
approximately 1 oz to 0.028 kg, or 1 kg to 35.2 oz. 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:
CR_ =
^
MS
(3-2)
where
Crmm
Cr,
lim ~
maximum allowable fish consumption rate (meals/mo)
maximum allowable fish consumption rate (kg/d)
3-5
-------
MS = meal size (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 14 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 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.
All meal consumption limits in the tables in Section 4 have been rounded down to
the nearest whole number of meals per month, with the exception of the 0.5-
meal/mo consumption limits, which are expressed as 6 meals/yr. Meal
consumption limits are rounded down to make them more protective;
rounding up would potentially cause them to exceed maximum ARLs for
carcinogens.
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 are 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
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 consump-
tion limits for chlordane for both carcinogenic and chronic (noncarcinogenic)
health effects are used to illustrate various modifications to the monthly con-
sumption limit tables. For developmental effects in women of reproductive age,
examples use PCBs as the target analyte.
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
3-6
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3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
Table 3-2. Input Parameters for Use in Risk Equations
Equation Parameter9
Maximum acceptable risk level (ARL)
Cancer slope factor (q-(*)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 (Tap)
Values
10"4 (unitless)
10'5 (unitless)
10'6 (unitless)
(mg/kg/d)'1
mq/kq/d
70 kg (general adult population)
70 kg (women of reproductive age)
14.5 kg (young children < 6 years)
3 oz (0.085 kg) (children only)
4 oz (0.1 14 kg)
8 oz (0.227 kg)
12oz(0.341 kg)
16 oz (0.454 kg) (adults only)
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)
14 d/14-d period (biweekly limit)
10 d/ 10-d period (10-day limit)
Selection of the appropriate maximum acceptable risk level, consumer body weight, and average fish meal size
are considered risk management decisions. This document provides a range of values for use in risk equations;
the risk management decision-making process is discussed in Volume 3 of this guidance series. Selection or
calculation of the appropriate cancer slope factor and RfD values may be considered a lexicological, medical,
or risk management decision. For information regarding these values, see Sections 2 and 5 of this document
and Volume 3.
Most of the q/s and RfDs were obtained from EPA's Integrated Risk Information System (IRIS, 1997). The
RfDs not listed in IRIS were obtained from EPA's Office of Pesticide Programs. The a *s 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.
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. Consumption limit tables for women of reproductive age
and children were calculated using RfDs for developmental effects for two target
analytes—methylmercury and PCBs. Section 5 contains a summary of the overall
toxicity of each target analyte. In future revisions of this document, additional
consumption limit tables for women of reproductive age and children based on
developmental health endpoints will be provided. The equations provided in this
3-7
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3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
IP
•So.
If!?ll
Ł
co
<" i;
-------
3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
section may be used by readers to calculate additional or alternative consumption
limits based on their interpretation of lexicological and other literature.
3.2.2.1 Calculation of Daily Consumption Limits-
Equation 3-3 is used to calculate the allowable daily consumption (CRlim) of
contaminated fish, based on a contaminant's noncarcinogenic health effects, and
is expressed in terms of kilograms of fish per day:
RfD • BW
(3-3)
where
CR!im =
RfE =
BW =
Cm =
maximum allowable fish consumption rate (kg/d)
reference dose (mg/kg/d)
consumer body weight (kg)
measured concentration of chemical contaminant m in a given
species of fish (mg/kg).
9Riim 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. This may, however, not be the
case for developmental toxicants (see Section 3.2.3).
In 1996, more than 75 percent of all fish consumption advisories issued in the
United States were issued wholly or in part as a result of methylmercury
contamination. Because the greatest exposure to methylmercury for most
individuals is related to consumption of fish and other seafoods, the Office of
Water believes a special discussion of methylmercury is warranted. In 1985, EPA
published in IRIS an oral RfD for methylmercury of 3 x 10'4 mg/kg/d for chronic
systemic effects for the general population. This was based on neurological effects
observed in adults in Iraq who ate methylmercury-contaminated grain. Because
of concern that the fetus may be at increased risk of adverse neurological effects
from exposure to methylmercury, EPA agreed to revise the RfD. In 1993, Volume
1 of this guidance series was published, and EPA's Office of Water recommended
use of an interim RfD of 6 x1(T5 mg/kg/d to protect against developmental effects.
This value was calculated by dividing the existing RfD of 3 x 10"4 mg/kg/d by a
factor of 5. This approach was to be considered interim until such time as the
Agency reevaluated new studies on the chronic and developmental effects of
methylmercury (U.S. EPA, 1995).
3-9
-------
5X10-"
1X1Q-3
5x10
I-* —
1x10"
5x10*-
1X101"
In May 1995, a new RfD of 1 x 1CT4mg/kg/d became available in IRIS. This was
based on neurological effects seen in Iraqi children who had been exposed to
methylmercury in utero. Data from other human studies also support this RfD. It
is acknowledged that the RfD is an estimate with uncertainty spanning perhaps an
order of magnitude. Sources of uncertainty include recording and classifying
events in the children's lives; uncertainty and variability in the pharmacologic
factors that were used in estimating the ingested mercury dose; and the relatively
small size of the study population (81 mother-child pairs). Despite these
uncertainties, this RfD value is a confident estimate of a level of exposure that can
be consumed on a daily basis without expectation of adverse health effects. Figure
3-1 shows the current RfD of 1 x 10'4 mg/kg/d and the upper and lower bounds of
the order of magnitude range around this RfD. The shaded area covers the factor
Fish tissue
concentrations*
(ppm)
Benchmark dose used in calculation of the current RfD
Current IRIS RfD value based p^dfvelopmental effects
Order of magnitude range of uncertainty
•Calculated assuming consumption of one 8-oz meal (uncooked weight) per week by a 70-kg adult.
Figure 3-1. Current RfD for Methylmercury with Associated Range of Uncertainty and
Corresponding Fish Tissue Contamination Levels.
3-10
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3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
of 3 on either side of the point estimate of the RfD contributing to the total order of
magnitude RfD range. As can be seen in Figure 3-1, the former RfD of 3 x 10'4 lies
at the top of the range, and the interim value of 7 x 10"5 recommended by the
Office of Water for developmental effects lies just below the lower end of the
range. The current RfD of 1 x 10'4 mg/kg/d is recommended in this document as
being protective against adverse health effects in all population groups—adults,
women of reproductive age, and children.
To calculate weekly fish meal consumption limits, Equation 3-3 was modified as
follows:
Cm ~
RfD x BW
CRr
lim
(3-4)
Using this equation, one can calculate the level of chemical contamination (Cm) in
a given species of fish assuming that a 70-kg adult consumes a maximum of one
8-oz (0.227-kg) meal/wk. As shown in Figure 3-1, the fish tissue concentration of
methylmercury, calculated using the current RfD value (1 x 10'4 mg/kg/d or 4.9 x
10" mg/wk) and the assumptions as stated above, gives a fish contamination
concentration of 0.22 ppm. Using the upper and lower bounds of the RfD range,
the corresponding fish tissue concentrations are 0.66 ppm and 0.15 ppm,
respectively. One reasonable interpretation of this analysis is that a 70-kg person
consuming one meal per week of fish contaminated with a range of 0.15 to 0.66
ppm methylmercury would be exposed to levels within the uncertainty range of the
RfD. The RfD is a risk assessment tool that describes an exposure that is
expected to be without appreciable risk. The data for methylmercury do not
support the prediction of risk for levels of exposure above the RfD. Shown in
Figure 3-1 is the benchmark dose, which served as the basis for the
methylmercury RfD. The benchmark dose here is the lower bound on an estimated
10 percent risk level for all effects observed in the Iraqi children exposed in utero.
In a large population exposed to the benchmark dose of 1 x 10"3 mg/kg/d, some
individuals would be expected to show effects.
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. Five meal
sizes were used to develop the consumption limit tables for noncarcinogenic
effects: 3 oz (children only), 4 oz, 8 oz, 12 oz, and 16 oz (adults only). 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
typically pertain to seasonal and subsistence fish consumers (see Section 2.4.5.4).
Where appropriate, risk assessors may choose to derive consumption limits based
on a shorter time-averaging period such as a 7-day, 10-day, or 14-day period (see
Section 3.3.6). These shorter time-averaging periods may be more appropriate to
apply to consumption scenarios for recreational fishers who fish primarily during
3-11
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3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
a 1- to 2-week vacation period. Note that, irrespective of the time-averaging period
selected (e.g., 7-day, 10-day, 14-day, 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. For
example, consuming the monthly consumption limit in 1 day (at one meal) could
result in exposures for noncarcinogens that are 30 times the RfD (see Section 2.3
for further discussion).
EXAMPLE 2: Calculating Monthly Consumption Limits for Chronic
Systemic Health Endpoints in the General Population for
Chlordane
Using Equations 3-3 and 3-2, the monthly meal consumption limits were
calculated for the chronic (noncarcinogenic) health effects of chlordane for
adults in the general population as shown in Table 3-4. Note: In comparing the
consumption limit tables for chlordane based on carcinogenic (Table 3-3) 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 (used to calculate screening values in Volume 1) 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.
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 carcino-
genic effects (see Table 3-2).
3,2.3 Calculation of Consumption Limits, for Developmental Effects
This guidance document considers three groups of fishers in deriving risk-based
consumption limits: adults in the general population, women of reproductive age,
and young children. Both women of reproductive age and young children are
known to be at risk from developmental toxicants. There is currently very limited
information on the potential reproductive and developmental impact due to
3-12
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Table 3-4. Monthly Consumption Limits for Chronic Systemic Health Endpoints for the
General Population—Chlordane
Concentration
in Fish Tissue9
(mg/kg or ppm)
0.01
0.02
0.03
0.04
0.05
0.06
0.07
0.08
0.09
0.1
0.2
0.3
0.4
0.5
0.6
0.7
0.8
0.9
1
2
>2
Recommended Risk— Based Consumption Limit (meals per month)"
4-oz Meal Size
(0.11 4 ka)
>30
>30
>30
28
22
18
16
14
12
1
5
3
2
. 2
6/yr
NONE
8-oz Meal Size
(0.227 ka)
>30
>30
28
18
14
11
9
8
7
6
5
2
1
1
1
6/yr
6/yr
6/yr
6/yr
6/yr
NONE
NONE
12-oz Meal Size
>30
>30
18
12
9
7
6
5
4
4
3
1
1
6/yr
6/yr
6/yr
6/yr
NONE
NONE
NONE
NONE
16-oz Meal Size
>30
28
14
9
7
5
4
4
3
3
2
1
6/yr
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE = No consumption recommended.
6/yr = Consumption of no more than 6 meals per year is recommended.
> 30 = Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 meal per day).
a Instructions for modifying the variables in this table are found in Section 3.3.
Consumption limits are calculated based on an adult body weight of 70 kg and using RfD = 6 x 10'5 mg/kg/d
References for RfDs are found in Section 5.
The detection limit is 1 x 10"3 mg/kg.
All values were rounded down to the nearest whole meal size
Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD).
When this dose is consumed in less than 1 month (e.g., in a few large meals), the daily dose will
exceed the RfD (see Section 2.3).
3-13
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damaged germ cells resulting from exposure of men to developmental toxicants.
However, some information on this is provided in Sections 2.3.2.3 and 5. It is well-
documented that women exposed to developmental toxicants (e.g., methylmercury
and PCBs) may transfer sufficient concentrations of these contaminants in utero
or through breast feeding to induce pre- or postnatal developmental damage in
their offspring. Information on developmental toxicity has only recently become
available for many chemical contaminants. Consequently, the IRIS database, the
source of most of the chronic RfDs used to calculate consumption limits, does not
contain sufficiently detailed developmental toxicity data for all target analytes.
Consumption limits specifically designed to address developmental health effects
in women of reproductive age and children have been calculated for
methylmercury and PCBs (see Section 4). For methylmercury, the RfD used in this
document (1 x 10"4 mg/kg/d) is based on developmental effects (see Section 3.2.2
for a detailed discussion).
For PCBs, the RfD used in this document (2 x 10'5 mg/kg/d) is based on the data
available for Aroclor 1254 for ocular and immunological effects in monkeys. The
RfD for Aroclor 1016 (7 x 1CT5 mg/kg/d) is based on developmental effects in
monkeys; however, this value is less conservative than the value for Aroclor
1254. Thus, EPA believes that the RfD based on Aroclor 1254 is protective for
adults, as well as for developmental effects in women of reproductive age and
children, and is using this value for all these groups in this document.
New developmental toxicity data on other target analytes are being reviewed by
EPA and will be incorporated in future editions of this document. Readers are
referred to Sections 2 and 5, which discuss developmental toxicity. Developmental
study data and sources of additional toxicity data are provided in Section 5 so that
readers may evaluate available information and make informed decisions con-
cerning developmental toxicity and consumption limits. Exposure limits for devel-
opmental and other health endpoints may be calculated by readers, as deemed
necessary, using methods described in Section 2.3.2.3.
EXAMPLE 3: Calculating Monthly Consumption Limits for Developmental
Endpoints in Women of Reproductive Age for PCBs
In calculating consumption limits for developmental effects in women of repro-
ductive age, Equation 3-3 is used to calculate the allowable daily consumption
rate (Cr,im) of contaminated fish (kilograms of fish per day), based on the RfD
of 7 x 10"5 mg/kg/d for Aroclor 1016, a bodyweight of 70 kg, and a range of
contaminant concentrations in fish tissue. Equation 3-2 is then used to calculate
the monthly consumption limits for a specified meal size. This developmental
Crr represents the maximum lifetime daily consumption rate (kg of fish per day)
thaTwould not be expected to cause adverse developmental effects.
(continued)
3-14
-------
3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
EXAMPLE 3 (continued)
The use of the RfD for developmental effects of PCBs is derived from the RfD
for Aroclor 1016 for which developmental effects have been demonstrated (see
Section 5.7 for a detailed discussion of toxicological data on PCBs). When
monthly meal consumption limits based on the developmental RfD (Table 3-5)
are compared to those for chronic (nondevelopmental) health effects that are
based on Aroclor 1254 for the general population (Table 3-6), the monthly meal
consumption limits based on nondevelopmental effects are more conservative
(restrictive) than those calculated using the developmental 'RfD. Risk assessors
and risk managers may wish to use the more restrictive meal consumption limits
developed using the RfD for Aroclor 1254 to be protective of both chronic
systemic and developmental health effects based on their population of
concern.
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. These values may not be valid
for some recreational and subsistence fisher populations. Readers are encouraged
to modify these variables, if necessary, to better fit local conditions relating to
consumer body weight, risk level of concern, meal size, new toxicological data,
and/or the presence of multiple contaminants. 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 (q.,*)
Chronic reference dose (RfD)
Consumer body weight (BW)
Fish meal size (MS)
Contaminant concentration in edible fish tissue (CJ
Time-averaging period (e.g., 7-day, 10-day, 14-day, and 30-day period).
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. Blank table templates are
provided in Appendix G for use in developing new consumption limit tables based
on changes in these input parameters.
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 ARLs of 10"4to 10"6. Note
3-15
-------
3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
Table 3-5. Monthly Consumption Limits for Developmental Health Endpoints for Women of
Reproductive Age—PCBs (Aroclor 1016)
Chemical
Concentration
in Fish Tissue*
(mg/kg or ppm)
<0.02
0.02
0.03
0.04
0.05
0.06
0.07
0.08
0.09
0.1
0.2
0.3
0.4
0.5
0.6
0.7
0.8
0.9
1
2
>2
Recommended Risk— Based Consumption Limit (meals per month)"
4-oz Meal Size
>30
>30
>30
>30
26
21
18
16
14
13
6
4
3
2
2
1
1
1
1
6/vr
NONE
8-oz Meal Size
(0.227 kcrt
>30
>30
21
16
13
10
9
8
7
6
3
2
1
1
1
6/yr
6/yr
6/yr
6/yr
NONE
NONE
12-oz Meal Size
(0.341 kq)
21
14
10
8
7
6
5
4
4
2
1
1
6/yr
6/yr
6/yr
6/yr
NONE
NONE
NONE
NONE
16-oz Meal Size
(0.454 ka)
16
„-> 10" ,„ t"
8. ^"
6"
5
4
4
3
3
1
1
6/yr
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE » No consumption recommended.
6/yr = Consumption of no more than 6 meals per year is recommended.
> 30 = Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 meal per day).
'Instructions for modifying the variables in this table are found in Section 3.3.
Consumption limits are calculated based on an adult body weight of 70 kg and using Rf D = 7 x 10'5 mg/kg/d.
References for RfDs are found in Section 5.
Shaded values are below the detection limit, 5 x 10"2 mg/kg.
All values were rounded down to the nearest whole meal size.
"Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD).
When this dose is consumed in less than 1 month (e.g., in a few large meals), the daily dose will
exceed the RfD (see Section 2.3).
3-16
-------
3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
Table 3-6. Monthly Consumption Limits for Chronic Systemic Health Endpoints for the
General Population—PCBs (Aroclor 1254)
Chemical
Concentration
in Fish Tissue9
(mg/kg or ppm)
0.006 ~"
0.007^
0.008
0.009 "
0.01 "
0.02
0.03
0.04
0.05
0.06
0.07
0.08
0.09
0.1
0.2
0.3
0.4
0.5
0.6
0.7
>0.7
Recommended Risk— Based Consumption Limit (meals per month)"
4-oz Meal Size
(0.1 14 kg)
>30
,>30
>30
>30
>30
>30 '
>30
>30
' 1-8
12
9
7 - - -
6
5
4
4
3
1
1
6/yr
6/yr
6/yr .
6/yr
NONE
8-oz Meal Size
(0.227 ka)
' -^ -^>30
>30 .',„«.
>30 , '"
>30
26
23
20
18
9
6
4
3
3
2
2
2
1
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
1 2-oz Mea! Size
(0.341 kq)
' >'30 •*"-*
>30 -
-- - 24"-
20 '-
17
> 15
• ' 13
12
6 -
4
3
'"•*•
2
1
1
1
1
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
16-oz Meal Size
(0.454 kg)
--• . *>30x, """*
" -.23 '"• «,
-; fa -
, 15
13 . .
.11 -*-
- fa
", 9 -. '=
• 4,
3- \
. , - 2
1 -. > \
1
1
1
1
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE = No consumption recommended.
6/yr = Consumption of no more than 6 meals per year is recommended.
> 30 = Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 meal per day).
a Instructions for modifying the variables in this table are found in Section 3.3.
Consumption limits are calculated based on an adult body weight of 70 kg and using RfD = 2 x 10"5 mg/kg/d.
References for RfDs are found in Section 5.
3-17
-------
3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
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 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. For
example:
• To go from an ARL of 10"4 to an ARL of 10"3, shift the meal consumption limit
values for the ARL of 10"4 to chemical concentrations 10 times greater than
the original (e.g., if the original meal consumption limit is 12 8-oz meals per
month at a contaminant concentration of 0.06 mg/kg for an ARL of 10~4, then
the new meal consumption limit for an ARL of 10'5 would be 12 8-oz meals per
month at a contaminant concentration of 0.6 mg/kg. This increases the
estimated individual lifetime risk by a factor of 10.
• To go from an ARL of 10~6 to an ARL of 10"7, shift the meal consumption
' values for the ARL of 10~6 to chemical concentrations 10 times lower than the
original (e.g., if the original meal consumption limit is 12 8-oz meals per month
at a contaminant concentration of 0.0006 mg/kg for an ARL of 10 , then the
new meal consumption limit for an ARL of 10"rwould be 12 8-oz rneals per
month at a contaminant concentration of 0.00006 mg/kg. This decreases the
estimated individual lifetime risk by a factor of 10.
EXAMPLE 4: Modifying the Acceptable Risk Level
The consumption limit table for adults in the general population for the carcino-
genic effects of chlordane was taken from Section 4 and modified to include
consumption limits for risk levels ranging from 10"3 (1 in 1,000) to 10"7 (1 in
10,000,000) for an 8-oz meal size as shown in Table 3-7. Consumption limits for
other meal sizes (4,12, and 16 oz) have been deleted from the table for clarity.
Using this table, the reader can easily see the tenfold effect of modifying the
ARL value on the monthly consumption limits for chlordane at specified con-
taminant concentrations in fish tissues.
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 q/s and RfDs were obtained from EPA
databases, primarily from IRIS (1997). 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).
EPA evaluates dose-response data as they become available. Because
toxicological data are continually being generated, there may be data available that
have not yet been incorporated into the risk values. This is especially relevant for
3-18
-------
3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
Table 3-7. Monthly Consumption Limits for Carcinogenic Health Endpoints
for the General Population—Chlordane
Chemical
Concentration
in Fish Tissue3
(mg/kg or ppm)
15
Recommended Risk — Based Consumption Limit (meals per month)b
4-oz
ARL
-
8-oz Meal Size (0.227 kg)
ARL
10'3
>30
>30
>30
>30
>30 "
>30
>30
>30
>30
- >30 '
>30
>30
>30
>30
>30
>30
>30
>30
>30
>30
>30
18
12
9
7
3
1
1
6/yr
6/yr
NONE
KT4
">30
,->30 >
530.,'-
>30 -,
>30
>30
>30
>30
>30
>30
>30
>30
>30
>30
>30
>30
18
12
9
7
3
1
1
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
ID"5
>30
>30
", ->30' •
-x >30""
->30 ,
>30
>30
>30 -
>30 -
>30
>30
18
12
9
7
3
1
1
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
10-*
>30
>30
>3Q
' >30
>30 -
'" >3Q "
1 18
J2
_9
7 -
3
1
1
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
ir/7
>30"~
1& "
12
9
T '
~ -3. ;
"i /" IS
'"' 1*"
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
12-OZ
ARL
-
- " -, --.
•*:,.
, ' ^^
~> ~ *
*" Ł
t -*•
,--
^~
-" "^
* ,ft
16-02
ARL
*: - =-.
, ~~ ~^
„, _^ -^
. •-' - -
~ X*
- •*•
". _~ -
---
:.... ** -- js
ARL = Acceptable Risk Level.
NONE = No consumption recommended.
6/yr = Consumption of no more than 6 meals per year is recommended.
> 30 = Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 meal per day).
a Shaded values are below detection limit, 1 x 10"3 mg/kg.. .
b Instructions for modifying the variables in this table are found in Section 3.3.
Consumption limits are calculated based on an adult body weight of 70 kg
and using a cancer potency factor of 1.3 per mg/kg/d.
References for cancer potency factors are found in Section 5.
All values were rounded down to the nearest whole meal size.
3-19
-------
3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
developmental toxicity, neurotoxicity, and imraunotoxicity data, which are the
subject of much current research. To address this, a summary discussion of the
toxicity data has been provided in Section 5 for each target analyte; the toxic-
ological profile summaries include a discussion of acute, chronic, developmental,
carcinogenic, and genetic toxicity, and special susceptibilities. The summaries
contain a brief synopsis of current toxicity data, based on a review of summary
documents and databases. Readers are urged to review this information for those
target analytes of interest in their geographic areas. A method for estimating
acceptable exposure levels from the new data is provided in Section 5; the
methods follow the basic dose-response approach recommended by EPA for
calculation of RfDs (discussed in Sections 2 and 5). This information is provided
to enable readers to use new dose-response data to develop consumption limits
if they feel the data warrant such an approach. This may be particularly useful
when there are concerns regarding the exposure of children and pregnant and
lactating women to developmental contaminants.
3.3.3 Consumer Body Weight (BW)
The consumption limit tables in Section 4 are based on fish consumers of two body
weights:
• 70 kg (156 Ib), the average body weight of male and female adults in the U.S.
population (U.S. EPA, 1990a).
• 14.5 kg (32 Ib), the body weight of a young child of 3 to 4 years in the U.S.
population (U.S. EPA, 1990a).
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 3-8 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. The reader should note that, because values in the tables in Section 4
were rounded down to the nearest number of whole meals, minor differences in
consumption limits for a specified meal size may occur when the reader calculates
the consumption limit directly using Equation 3-2 with the specified body weight
found in Table 3-8.
3-20
-------
3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION UM(TS
Table 3-8. Average Body Weights and Associated Multipliers
Age Group
(yr)a
<3
3 to 6
Oto6
6 to 9
9 to 12
12 to 15
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
Average Male
Body Weight (kg)
11.9
17.6
14.8
25.3
35.7
50.5
64.9
73.7
78.7
80.8
81.0
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
67.9
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
—
71.8(70)°
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 70-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.
0 Per recommendations in the Exposure Factors Handbook, the body weight value of 71.8 kg for the general adult
population was rounded to 70 kg (U.S. EPA, 1990a).
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:
Multiplier = Altemative Consumer Body Weight
BW General Adult Body Weight
(3-5)
3-21
-------
3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
EXAMPLE 5: Calculating Multiplier for Alternative Body Weight
To calculate the multiplier for a 9- to 12-year-old child with an average body
weight of 36 kg (see Table 3-8), the equation is:
MultiplierBW =
70 kg
= 0.51 (unitiess)
where 0.51 represents the ratio between the child body weight (36 kg) and the
adult body weight (70 kg). In this document, 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. A body weight of 14.5 kg was used in this document
to calculate all consumption limits for children.
Readers may wish to modify some or all of these consumption limits using
alternative body weight values, based on the health endpoint of concern. For
example, if certain developmental toxicants are of concern, exposure of women
of reproductive age could be assessed separately. As described in Section 2, a
body weight value of 64 kg (143 pounds) can be used to represent the body weight
for women of reproductive age, based on the arithmetic mean of the average
weights of women of three age groups (18-25, 26-35, and 36-45 years) given in
the Exposure Factors Handbook (U.S. EPA, 1990a; see Table 3-8). A more
protective body weight value to use would be the lower 95th percentile body weight
of women age 18 to 25 years (Blindauer, 1994). Readers are encouraged to use
local information when available to determine appropriate body weights for
calculating exposure limits for populations of concern.
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 population by the multiplier associated with the new
body weight:
New
MultiplierBW
(3-6)
where
Crmm = maximum allowable fish consumption rate (meals/mo)
mm
CRmm70_kg BW = maximum allowable fish consumption rate of a 70-kg fish
consumer (meals/mo)
BW = consumer body weight (kg)
MultiplierBW = body weight multiplier (unitless).
If the resultant meal consumption limit is not a whole number, round down to the
nearest whole number. Rounding up would make the meal consumption limit less
3-22
-------
3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
protective. It is also important to note that rounding down the modified CR,jm could
lead to unnecessarily conservative meal consumption limits, since the original
CR|jm values for adults have already been rounded down once. Readers may wish
to derive more accurate estimates by recalculating meal consumption limits using
Equations 3-1 through 3-3 instead of using a multiplier.
EXAMPLE 6: Modifying Consumption Limits for an Alternative Body
Weight
To modify Table 3-4 (which shows the monthly consumption limits for chronic
systemic health effects in the general adult population from exposure to
chlordane) to represent consumption limits for 9- to 12-year-old children
weighing an average of 36 kg, multiply each value in the consumption limit
table by 0.51:
New CR^ =
0.51
Table 3-4 has been modified to represent consumption limits for 36-kg
consumers, as shown in Table 3-9. For example, while adults may be able to
safely eat nine 8-oz fish meals contaminated with 0.06 mg chlordane/kg fish
per month, 9- to 12-year-old children would only be able to safely eat four 8-oz
fish meals per month at this same contaminant level (9 • 0.51 = 4.59, rounded
down to 4) to achieve the same daily contaminant exposure in mg/kg/d.
Scaling to a different body weight may cause the contaminant concentrations
of concern to change substantially. Note, for example, that in Table 3-9 all meal
consumption limits listed in the original table were effectively halved at any
given contaminant concentration to derive the new meal consumption limits for
36-kg consumers. Section 3.3.4 describes the methods used to add meal
consumption limits to concentration ranges that fall below the >30 meals per
month (">30") or above the no consumption ("NONE") designations in the
original table for the 70-kg consumer but no longer do in the modified table. For
the purposes of this document, safe fish consumption (represented by ">30"
meals in the tables) has been defined as an intake limit of 30 meals per month
and no more than one meal per day, for the monthly tables. The definition of
no safe fish consumption (represented by "NONE" in the tables) is a health
intake limit of less than one meal every 2 months for the monthly tables.
3.3.4 Meal Size
Meal size is defined as the amount of fish (in kilograms) consumed at one meal
Four average meal sizes (e.g., 4, 8, 12, and 16 oz) were used to develop
consumption limits for the general adult population. Four average meal sizes (e.g.,
3, 4, 8, and 12 oz) were also used for children under 4 years of age. Note:
Although children tend to eat smaller portions than adults, they may consume
significantly more fish per unit of body weight. Women of reproductive age
(considered separately for methylmercury and PCB consumption limits) were
assumed to eat the same amount of fish per meal as other adults.
3-23
-------
3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
Table 3-9. Monthly Consumption Limits for Chronic Systemic Health Endpoints for
Children—Chlordane
Chemical
Concentration
in Fish Tissue8
(mg/kg or ppm)
<0.005
0.005
0.006
0.007
0.008
0.009
0.01
0.02
0.03
0.04
0.05
0.06
0.07
0.08
0.09
0.1
0.2
0.3
0.4
0.5
0.6
0.7
0.8
0.9
1
1.1
>1 1
Recommended Risk— Based Consumption Limit (meals per month)"
4-oz Meal Size
(0114kq)
>30
>30
>30
>30
>30
>30
>30
28
19
14
11
9
8
7
6
5
2
1
1
1
6/yr
6/yr
6/yr
6/yr
6/yr
6/yr
NONE
8-oz Meal Size
(0.227 kq)
>30
>30
>30
>30
>30
>30
28
14
9
7
5
4
4
3
3
2
1
6/yr
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE
12-oz Meal Size
(0.341 kq)
>30
>30
>30
27
24
21
19
9
6
4
3
3
2
2
2
1
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
16-oz Meal Size
(0.454 ka)
>30
28
24
20
18
16
14
7
4
3
2
2
2
1
1
1
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE = No consumption recommended.
6/yr = Consumption of no more than 6 meals per year is recommended.
> 30 = Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 meal per day).
* Instructions for modifying the variables in this table are found in Section 3.3.
Consumption limits are calculated based on a body weight of 36 kg and using RfD = 6 x 10"5 mg/kg/d.
References for RfDs are found in Section 5.
The detection limit is 1 x 10"3 mg/kg.
All values were rounded down to the nearest whole meal size.
b Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD).
When this dose is consumed in less than 1 month (e.g., in a few large meals), the daily dose will
exceed the RfD (see Section 2.3).
3-24
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3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
EPA has identified a value of 8 oz (227 g) of cooked fish fillet per 70-kg consumer
body weight as an average meal size for adults in the general population and
women of reproductive age assuming consumption of noncommercially caught fish
only. This meal size, however, does not represent higher end exposures—persons
who consume more than the average in a given meal. These larger meal sizes are
particularly important to consider in cases where acute or developmental effects
from fish contamination are of concern. Therefore, the consumption limit tables in
Section 4 provide meal limits based on a range of meal sizes from 4 to 16 oz. For
children younger than 4 years of age, EPA has estimated the average fish meal
.size to be,3 oz (85 .g), of cooked fish fillet. Meal sizes of 3 to 12 oz are used in
Section 4 for,ehildren-under 4 years. Development of these values is discussed in
Section 2. . , . :„
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 D). Table 3-10 provides
alternative meal sizes and their associated multipliers. To obtain modified con-
Table 3-10. Alternative Meal Sizes and Associated Multipliers
Meal Size (oz)a
Multiplier11
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
24
32
8
4
2.67
2.0
1.6
1.33
1.14
1.0
0.89
0.80
0.73
0.67
0.62
0.57
0.53
0.5
0.33
0.25
* Bolded values are those used in the consumption limit tables in Section 4.
The meal size multiplier is multiplied by the consumption limits associated with 8-oz meals to
obtain new consumption limits using the alternative meal size.
3-25
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3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
sumption limits using alternative values for meal size, multiply the existing con-
sumption limits found in the tables for the 8-oz meal size by the multiplier
associated with the new meal size:
New CR^ =
-ozMS
Multiplier^
(3-7)
where variables are as previously defined. If the resultant consumption limit (meals
per month) is not a whole number, round down to the nearest whole number). For
the purposes of this document, safe fish consumption (represented by ">30" meals
in the tables) has been defined as an intake limit of 30 meals per month and no
more than one meal per day for the monthly tables. The definition of no safe fish
consumption (represented by "NONE" in the tables) is a health intake limit of less
than one meal every 2 months for the monthly tables.
EXAMPLE 7: Modifying Consumption Limits for an Alternative Meal Size
To modify Table 3-4 to develop values for a 24-oz meal size, multiply the
consumption limits for 8-oz meals in the table by one-third, or 0.33:
New CR^ =
0.33 .
Table 3-4 has been modified as shown in Table 3-1 1 to represent consumption
limits for adults consuming 24-oz (1-1/2-lb or 0.680-kg) fish meals. This shows
that, although a 70-kg adult could consume nine 8-oz fish meals contaminated
with 0.06 mg chlordane/kg fish, he or she could only consume three 24-oz
meals per month (9 • 1/3 = 3).
Scaling to a different meal size may cause the concentrations of concern to
change substantially. Note that in Table 3-1 1 , meal consumption limits listed in
the original table (Table 3-4) were divided by 3 to derive the new consumption
limits for 24-oz meals. Sections 3.2.1 .2 and 3.2.2.2 describe the methods used
to derive meal consumption limits (meals per month) for fish contaminant
concentration ranges that fall below the >30 meals per month (>30) or above the
no consumption ("NONE") range in the original table for the 8-oz meal size, but
no longer do in the modified table.
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
kg BW.S-oz MS
MultiplierBW • Multiplier^ (3_g)
where the parameters are as previously defined.
3-26
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3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
Table 3-11. Monthly Consumption Limits for Chronic Systemic Health Endpoints for the
General Population—Chlordane
Chemical
Concentration
in Fish Tissue8
(mg/kg or ppm)
<0.007
0.007
0.008
0.009
0.01
0.02
0.03
0.04
0.05
0.06
0.07
0.08
0.09
0.1
0.2
0.3
0.4
0.5
0.6
0.7
0.8
0.9
1
2
>2
Recommended Risk— Based Consumption Limit (meals per month)"
4-oz
Meal Size
(0.114kg)
>30
>30
>30
>30
>30
>30
>30
28
22
18
16
14
12
11
5
3
2
2
1
1
1
1
1
6/yr
NONE
8-oz
Meal Size
(0.227 kg)
>30
>30
>30
>30
>30
28
18
14
11
9
8
7
6
5
2
1
1
1
6/yr
6/yr
6/yr
6/yr
6/yr
NONE
NONE
12-oz
Meal Size
(0.341 kg)
>30
>30
>30
>30
>30
18
12
9
7
6
5
4
4
3
1
1
6/yr
6/yr
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
16-oz
Meal Size
(0.454 kg)
>30
>30
>30
>30
28
14
9
7
5
4
4
3
3
2
1
6/yr
6/yr
6/yr
NONE .
NONE
NONE
NONE
NONE
NONE
NONE
24-oz
Meal Size
(0.680 kg)
>30
26
23
20
18
9
6
4
3
3
2
2
2
1
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE = No consumption recommended.
6/yr = Consumption of no more than 6 meals per year is recommended.
> 30 = Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 meal per day).
a Instructions for modifying the variables in this table are found in Section 3.3.
Consumption limits are calculated based on adult body weight of 70 kg and using RfD=6x10"5 mg/kg/d.
References for RfDs are found in Section 5.
The detection limit is 1x10"3 mg/kg.
All values were rounded down to the nearest whole meal size.
b Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD).
When this dose is consumed in less than 1 month (e.g., in a few large meals), the daily dose will
exceed the RfD (see Section 2.3).
3-27
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3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
EXAMPLE 8: Modifying Consumption Limits for Alternative Body Weight
and Meal Size
To modify Table 3-4 for a 3-oz meal size for 3- to 6-year-olds with ah average
body weight of 17.4 kg, multiply the values for the 8-oz meal size by 0.25 (the
body weight multiplier; see Table 3-8 ) and 2.67 (the meal size multiplier; see
Table 3-10):
New
kg BW.8-OZ MS
0.25 • 2.67 .
In this example, the new consumption limits are equivalent to two-thirds (0.67)
of the old consumption limits. Table 3-4 has been modified to represent
consumption limits for 17.4-kg children consuming 3-oz (0.085 kg) fish meals,
as shown in Table 3-12. For clarity, only the meal frequencies for a 3-oz meal
size have been calculated. This shows that, although a 70-kg adult could
consume nine 8-oz fish meals contaminated with 0.06 mg/kg chlordane (see
Table 3-4), a 17.4-kg child could consume only six 3-oz meals per month at the
same fish contamination level (9 • 0.25 • 2.67 = 6). Again, readers will need to
use the methods described in Sections 3.2.1.2 and 3.2.2.2 to add consumption
limit values to concentration ranges that fall below 30 meals per month (>30) or
above the no consumption ("NONE") range in the original table for adults, but
no longer do in the modified table.
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, 1992), 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
calculation of the risk-based consumption limits should be derived from monitoring
data obtained from fish sampling and analysis programs and are specific to the
waterbody, fish species, and fish size (age) class that were sampled.
3-28
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3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
Table 3-12. Monthly Consumption Limits for Chronic Systemic Health Endpoints for
Children—Chlordane
Chemical
Concentration
in Fish Tissue8
(mg/kg or ppm)
<0.02
0.02
0.03
0.04
0.05
0.06
0.07
0.08
0.09
0.1
0.2
0.3
0.4
0.5
0.6
0.7
>0.7
Recommended Risk— Based Consumption Limit (meals per month)"
3-oz Meal Size
(0.085 kq)
>30
18
12
9
7
6
5
4
4
3
1
1
6/yr
6/yr
6/yr
6/yr
NONE
4-oz Meal Size
(0.11 4 kq)
8-oz Meal Size
(0.227 kq)
12-ozMeal Size
(0.341 kq)
NONE = No consumption recommended.
6/yr = Consumption of no more than 6 meals per year is recommended.
> 30 = Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 meal per day).
a Instructions for modifying the variables in this table are found in Section 3.3.
Consumption limits are calculated based on a body weight of 17.4 kg and using Rf D=6x10"5 mg/kg/d.
References for RfDs are found in Section 5.
The detection limit is 1 x 10'3 mg/kg.
All values were rounded down to the nearest whole meal size.
"Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD).
When this dose is consumed in less than 1 month (e.g., in a few large meals), the daily dose will
exceed the RfD (see Section 2.3).
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 for noncarcinogens 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-
3-29
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3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
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. Monthly consumption limits typically pertain to seasonal and sub-
sistence consumers who consume noncommercially caught fish as a major part
of their diets for a greater percentage of the year than recreational fishers. For
chemical contaminants with carcinogenic properties, there is no current method-
ology 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 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.
For chemical contaminants with noncarcinogenic effects, the situation is slightly
different especially when developmental effects are involved. Risk managers may
wish to calculate alternative consumption limits for different time intervals that
better represent specific populations. For example, 10-day consumption limits are
often calculated to apply to fish consumption scenarios for short-term recreational
fishers. The 10-day averaging period is a useful period for several reasons (see
Section 2.4.3.2). The 10-day averaging period is one of the averaging periods
used by the EPA Office of Water in developing health advisories for drinking water,
it is also relevant to the short time period often considered critical for exposure to
developmental toxicants, and this period corresponds to a typical vacation period.
Using chlordane as an example, 10-day fish consumption limits were calculated
for adults in the general population as shown in Example 9.
EXAMPLE 9: Calculating Wleal Consumption Limits for a Time-Averaging
Period of 10 Days
The monthly meal consumption limits shown in Table 3-4 include a limit for a
chlordane contaminant concentration of 0.07 mg/kg fish and a meal size of 4 oz
(0.114 kg), based on a chronic systemic RfD of 6 x 10"5 mg/kg/d and a con-
sumer body weight of 70 kg. Using these values for the parameters in Equation
3-3 yields a maximum daily consumption rate of 0.060 kg fish/d:
6 x, 10 5 mg/kg-d • 70 kg =
0.07 mg/kg
fish/d
(continued)
3-30
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3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
EXAMPLE 9 (continued)
Using a meal size of 0.114 kg (4 oz), a consumption rate of 0.060 kg fish/d, and
a time-averaging period of 10 days (expressed as 10 days/10-day period),
Equation 3-2 yields a meal consumption limit of five meals per 10-day period,
as shown in Table 3-13.
- 0.060 kg/d • 10 days/10-day period "-',„„.
-- - .. . , - - — =LJ- - =5 meals/10-day period.
0.114 kg/meal
All meal consumption limit modifications may be done in this manner by
substituting the appropriate values into Equations 3-1 through 3-3. Note that
while meal sizes are given in ounces in the tables in Section 4, they need to be
calculated in kilograms in Equations 3-1 through 3-3. The conversion rate from
ounces to kilograms is approximately 1 oz to 0.028 kg, or 1 kg to 35.2 oz.
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:
(3-9)
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 j 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
3-31
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3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
Table 3-13. 10-Day Consumption Limits for Chronic Systemic Health Endpoints
for the General Population—Chlordane
Chemical
Concentration
in Fish Tissue8
(mg/kg or ppm)
<0.009
0.009
0.01
0.02
0.03
0.04
0.05
0.06
0.07
0.08
0.09
0.1
0.2
0.3
Recommended Risk— Based Consumption Limit (meals per 10 days)b
4-oz Meal Size
(0.114kg)
>10
>10
>10
>10
>10
9
7
6
5
4
4
3
1
1
NONE
8-oz Meal Size
(0.227 kg)
>10
>10
>10
9
6
4
3
3
2
2
2
1
NONE
NONE
NONE
12-oz Meal Size
(0.341 kg)
>10
>10
6
4
3
2
2
1
1
1
1
NONE
NONE
NONE
16-oz Meal Size
(0.454 kg)
10
9
4
3
2
1
1
1
1
1
NONE
NONE
NONE
NONE
NONE = No consumption; less than 1 meal per 10 days.
> 10 - Although consumption of more than 10 meals in a 10-day period is allowed, EPA advises limiting
consumption to 10 meals in a 10-day period (1 meal per day).
* Instructions for modifying the variables in this table are found in Section 3.3.
Consumption limits are calculated based on adult body weight of 70 kg and using Rfp=6x10~5mg/kg/d.
References for RfDs are found in Section 5.
The detection limit is 1 x lO^mg/kg-d.
AH values were rounded down to the nearest whole meal size.
b 10-day limits are based on the total dose allowable over a 10-day period (based on the RfD).
When this dose is delivered in less than 10 days (e.g., in a single meal), note that the daily dose will
exceed the RfD (see Section 2.3).
(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 carcino-
genic 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
weighted average chemical contaminant concentration across all of the species
consumed:
3-32
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3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LflWfTS
ARL • BW
(3-10)
where
CRnm = 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)
P: = proportion of a given species in the diet (unitless)
q., = 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 = CRHm ' P,
(3-11)
where
CRlim -
j
= consumption rate of fish species j (kg/d)
maximum allowable fish consumption rate (kg/d)
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 CRnm 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 concen-
tration across all of the species consumed:
RfD • BW
Ł (cmj
(3-12)
mj
3-33
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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.
EXAMPLE 10: Calculating Consumption Limits for a 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 0.00006 mg/kg/d (IRIS, 1997). Since chlordane causes both chronic health
and carcinogenic effects, consumption limits must be calculated for both health
endpoints. The q.,* for chlordane reported in IRIS is 1.3 per (mg/kg/d)'1 (IRIS,
1997). 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 0.073 kg/d, based on
carcinogenic endpoints:
10 s • 70 kg
(0.006 mg/kg • 0.3 + 0.008 mg/kg • 0.7) • 1.3 per mg/kg/d
= 0.073 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):
= 0.073 kg/d • 30.44 d/mo = g g „ g meals/mo
0.227 kg/meal
Equation 3-2 yields a meal consumption limit of nine 8-oz meals per month,
based on chlordane's carcinogenicity. Equation 3-7 indicates that two 8-oz
catfish meals and six 8-oz trout meals per month may be consumed using an
acceptable cancer risk level of 10~5.
= 9 meals/mo ' 0.3 = 2.7 « 2 meals/mo
= 9 meals/mo • 0.7 = 6.3 « 6 meals/mo .
(continued)
3-34
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3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
EXAMPLE 10 (continued)
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. In this case, since catfish were less contaminated than
trout, the consumption limit would be three 8-oz catfish meals and six 8-oz trout
meals per month. This approach is based on the proportion of each type of fish
eaten. If the actual proportion differs from that used in the equation, the resulting
risks may be higher or lower than the targeted risk level (e.g., 10"^.
Noncarcinogenic Effects: Equation 3-8 is used to calculate the daily
consumption limit based on chlordane's chronic health effects using the RfD
rather than the q.,*.
6 x 10 s mg/kg/d • 70 kg
0.006 mg/kg • 0.3 + 0.008 mg/kg • 0.7
= 0.570 kg/d .
As with carcinogenic effects, Equation 3-2 is used to convert the daily consump-
tion limit of 0.570 kg fish to a meal consumption limit:
ittni
_ 0.570 kg/d • 30.44 d/mo
0.227 kg/meal
= 76.4 ~ 76 meals/mo
This analysis indicates that 0.57 kg/d is equivalent to 76 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 ">30" meals /month).
and has been defined as an intake limit of 30 meals per month and no more
than one meal per day 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 of two (or three) 8-oz catfish meals and six 8-oz
trout meals per month based on chlordane's carcinogenic effects, the more
sensitive of the two health endpoints, or elect to base their advisories on
noncarcinogenic effects. In addition, they may also review the toxicity data and
develop exposure limits based on their interpretation of the toxicity data.
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.
3-35
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3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
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.
Possible toxic interactions in mixtures of chemicals are usually placed in one of
three categories:
• 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, 1986d), 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 and Appendix C. For mixtures of chemicals that produce similar tox-
icological endpoints, consumption limits are derived by summing the contaminants
from all fish species consumed, as discussed below.
Some chemical mixtures may contain chemicals that produce dissimilar health
effects. Methods currently do not exist for combining dissimilar health effects to
characterize overall health concerns from chemical mixtures. Instead, the risks
from these contaminants need to be characterized and presented separately. It is
important to review the overall toxicity of a contaminant in evaluating combined
risks because most contaminants are capable of causing multiple effects in
numerous organ systems at elevated exposures.
3.5.1 Carcinogenic Effects
Cancer evaluations and q^s for most carcinogens are based on animal studies,
which are not generally assumed to predict the site of cancer in humans.
Consequently, carcinogenic effects are not usually categorized by the cancer site
observed in animal studies. Rather, carcinogenic effects are generally assumed
to be additive, unless data derived from human studies show them to be otherwise.
3-36
-------
3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
Carcinogenic effects for the 14 carcinogens considered in this guidance series are
assumed to be additive.
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 Equation
3-1, with the summation of all species and all chemicals substituted for Cm in the
denominator:
ARL • BW
(3-13)
m=lVj-l
where
ARL
BW
= maximum allowable fish consumption rate (kg/d)
= maximum acceptable lifetime risk level (unitless)
= consumer body weight (kg)
= concentration of chemical contaminant m in species j (mg/kg)
= proportion of a given species in the diet (unitless)
= 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:
m
ARL • BW
(3-14)
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
substituted for Cm in the denominator and their respective RfDs in the numerator:
m=l
RflX
E (cmj • pp
BW
(3-15)
3-37
-------
3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LIMITS
where the parameters are as previously defined. 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:
BW
(3-16)
m=l
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 70 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 q.,* for chlordane reported in IRIS is 1.3 per
(mg/kg/d) (IRIS, 1997). The q^ for heptachlor epoxide reported in IRIS is 9.1
per (mg/kg/d) (IRIS, 1997). Equation 3-10 is used to calculate daily consumption
rate based on the combined carcinogenic effects of both contaminants:
1(T5 • 70
(0.04 • 1.3) + (0.01 • 9.1)
= 0.005 kg/d .
A daily consumption rate of 0.005 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 heptachlor are 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, 1997). A combined daily consumption limit based
on an RfD of 6 x 10'5 mg/kg/d for chlordane and 1.3 x 10'5 mg/kg/d for hepta-
chlor was calculated using Equation 3-12:
(continued)
3-38
-------
3. DEVELOPMENT AND USE OF RISK-BASED CONSUMPTION LflWfTS
EXAMPLE 11 (continued)
Equation 3-12 yields a daily consumption rate of 0.2 kg fish/d, at the con-
taminant concentrations described above. Using Equation 3-2, a meal consump-
tion limit of 26 (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 con-
sumption to 26 4-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.
Risk assessors or risk managers may, however, elect to use a third health
endpoint (e.g., developmental toxicity), based on their review of the toxicological
data for these chemicals.
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. CRnm 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 reported in IRIS is 0.003 mg/kg/d (IRIS, 1997), and the RfD for diazinon
is 0.00009 mg/kg/d (IRIS, 1997; U.S. EPA, 1993d).
' • (continued)
3-39
-------
EXAMPLE 12 (continued)
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 0.060 kg fish per
day is calculated using Equation 3-11:
0.003
0.00009
70
(4.0 • 0.7) + (6.0 • 0.3) (0.3 • 0.7) + (0.8 • 0.3),
= 0.060 kg/d .
Using Equation 3-2, a meal consumption limit of eight 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).
Equation 3-7 is used to determine meal consumption limits for trout and catfish,
based on a diet of 70 percent trout and 30 percent catfish:
= 8 8-oz meals/mo • 0.7 = 5.6 » 5 8-oz meals/mo .
According to Equation 3-7, an adult may safely consume five 8-oz meals of trout
and two 8-oz meals of catfish per month. Again, as mentioned in Section 3.4.2,
rounding down both species-specific consumption limits 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
offish meals per, month .equals that found in Equation 3-11. In this case, since
trout were less contaminated than catfish,-the' species-specific consumption limit
would be six 8roz trout meals and two 8-oz catfish meals per month.
3.6 CHOICE OF CONSUMPTION LIMITS
Where chemicals found in a given species cause dissimilar health effects, dose
addition is not often justified scientifically; if dose addition is used in these cases,
it must be supported by biological plausibility (U.S. EPA, 1986d). Thus, in most
cases where chemicals cause different effects, readers are advised to use the
consumption limit for the contaminant that results in the most protective fish
consumption advisory for the population of concern. This approach may result in
different advisories for each population of concern, since different populations may
3-40
-------
have varying health effects of concern (e.g., developmental effects for women of
reproductive age and children vs. other chronic effects for adults in the general
population). It is important to compare the consumption limits calculated for all
chemicals found in a given species and choose the most appropriate consumption
limits for each consumer population. This selection may be considered a risk
management decision or a medical/toxicological decision.
If local fish sampling and analysis programs and fish consumption surveys
determine that consumers are exposed to all four of the contaminants discussed
in the above examples (chlordane, heptachlor epoxide, chlorpyrifos, and diazinon)
then risk assessors could consider a number of separate fish consumption limits
for the chemical mixture:
• A consumption limit could be based on the combined carcinogenic effects of
chlordane and heptachlor epoxide, as shown in Example 11.
• A consumption limit could be based on chronic liver damage caused by
exposure to the chlordane and heptachlor, as shown in Example 11.
• A consumption limit could be developed for the chronic health effect
(cholinesterase inhibition) caused by exposure to both chlorpyrifos and
diazinon, as shown in Example 12.
• Consumption limits could be based on new study data reviewed by readers
(e.g., xenoestrogenic effects of organochlorines or developmental effects)
using the equations provided in this document. Intake limits based on
developmental or other toxicological effects can be calculated using the same
equations as are used for chronic health endpoints throughout Section 2.
Readers might then base fish consumption advisories for this particular chemical
mixture on the most appropriate of the consumption limits for each population of
concern. Decisions regarding selection and implementation of consumption limits
will be discussed in Volume 3 of this series.
Most fish contaminants have the potential to cause multiple adverse health effects
given sufficiently high exposure levels. Readers are advised to examine the
toxicological profile summaries in Section 5 to determine the most appropriate
health endpoints on which to base fish consumption limits. Many of the chronic
RfDs for organophosphates are currently based on cholinesterase inhibition as the
critical health endpoint of concern. However, the EPA's Scientific Advisory Board
recently raised concerns about using cholinesterase inhibition as a critical endpoint
in the absence of clinical symptoms (U.S. EPA, 1993q). For those who wish to
calculate alternative exposure limits for these or other chemical contaminants for
use in developing alternative consumption limits, Section 2 provides a summary
of current EPA methods for doing so and sources of additional guidance.
3-41
-------
-------
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 and chronic health endpoints for young
children (<6 years) for 23 of the target analytes listed in Table 1-1. In addition,
consumption limit tables based on developmental effects associated with exposure
to methylmercury and PCBs are included for women of reproductive age and
children. The consumption limits in these tables were calculated using Equations
3-1 through 3-3 and the default values for each of the variables described in
Sections 3.2 and 3.3.
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 (q/) and the maximum acceptable
risk level for carcinogenic health endpoints. Three maximum ARLs are offered as
options in the consumption limit tables for carcinogenic effects. Note: Selection of
the most appropriate risk level is a risk management decision to be made at the
State, local, or Tribal organization level and is discussed in Volume 3. Default
values for the other variables are presented in Section 3 and described in greater
detail in Section 2.
Current EPA risk values (cancer slope factors and RfDs) were used in the
consumption limit calculations, as described in Section 3.2 and Section 5. Because
there are new toxicity data for some target analytes not reflected in the current
RfDs, readers are encouraged to review the toxicity data in the toxicological profile
summaries given in Section 5. Methods for using this information to generate
additional consumption limits are discussed in Section 2.3.
Most target analytes are known to cause multiple effects at elevated exposure
levels, according to the results of animal studies. However, because the levels of
chemical contaminants in fish typically ingested by consumers are lower than
those used in animal studies, consumers may experience only one or a few of the
known chronic effects, and/or only the more susceptible members of the
population may develop the health endpoints of concern. For example, analytes
that cause developmental toxicity pose a risk to children, pregnant women, nursing
mothers, and women of reproductive age who are planning to have children.
4-1
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Readers are advised to consider these data in reviewing the consumption limit
tables and the toxicity discussions in Section 5. Alternatively, some effects are of
concern to all exposed individuals. For example, women of reproductive age do
not necessarily develop cancer or chronic exposure health effects at different rates
than others in the general population. Consequently, the consumption limit tables
for carcinogenic and chronic exposure toxicity for the general population are also
applicable to this population of concern. Note: Women of reproductive age are
defined as a separate population only for health intake limits based on develop-
mental health effects.
Monthly consumption limits were developed for carcinogenic health endpoints and
for chronic exposure health endpoints. Time-averaging periods of less than 1
month (e.g., 7-day, 10-day, or 14-day periods) can be used by risk assessors or
managers to develop consumption limits for chronic exposure health endpoints as
described in Section 3.3.6. Note: Both the monthly consumption limit tables
(shown in this section) and short-term consumption limit tables '(e.g., 7-day, 10-
day, and 14-day) that may be developed by risk managers are based on the same
chronic RfDs because RfDs for acute health endpoints are currently not available
from the EPA RfD/RfC Workgroup (see Section 5). The use of these short-term
time-averaging periods applies to short-term recreational fishers (e.g., vaca-
tioners), and monthly consumption limits pertain to both seasonal (e.g., 3-month)
and year-round subsistence fishers.
Monthly consumption limit tables were developed for chronic systemic health
effects for adults and young children for 23 and 22 of the 25 target analytes,
respectively. Monthly consumption limit tables using reference doses based on
developmental endpoints were calculated for both children and women of
reproductive age for methylmercury only. Readers may wish to calculate
consumption limits for young children based on developmental toxicity data
discussed in Section 5 for some of the other target analytes as has been done for
methylmercury. Section 2.3 describes a methodology that can be used to derive
risk-based consumption limits for developmental and other toxicities.
Each consumption table lists, by chemical and fisher population, the maximum
number of fish meals per unit time (monthly) that may be safely eaten by the
population of concern. Limits are given for a range of fish meal sizes, population
members (adults, women of reproductive age, and children), and chemical
contaminant levels in fish tissues. In addition, for carcinogens, limits are provided
for a range of maximum ARLs (10:4 to 1Q-6). Readers may use these tables by: (1)
determining the chemical contaminant concentration in fish surveyed in local fish
sampling and analysis programs, (2) determining the meal size (and risk level, for
carcinogens) they wish to use, (3) locating the point where these parameters
intersect on the consumption limit table, and (4) reading the value for the maximum
number of meals per month that may be safely eaten for each contaminant. All
consumption limits have been rounded down to the nearest whole number of meal
per month, with the exception of consumption limits of 0.5 meal per month, which
are expressed as 6 meals/year in the tables.
4-2
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
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 (LOD) are provided so that risk managers may use them once
lower detection limits are achievable through improvements in analytical
procedures . Current limits of detection are shown in the footnotes at the bottom
of each of the consumption limit tables arid additional information is provided in
Appendix H. In the consumption tables, all fish tissue concentrations below the
detection limits and the associated number of meals per month are shaded for
emphasis. 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. The reader 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 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.
For the purposes of this document, safe fish consumption (represented by
">30" in the tables) has been defined as a consumption limit of 30 fish meals
per month assuming nb more than one meal per day) for the monthly
consumption limit tables. In the case of the monthly consumption limits, where
risk calculations indicate that more than 30 meals per month may be eaten without
incurring undue risk of cancer or developing adverse chronic health effects, risk
managers may choose not to issue a fish consumption advisory for the population
of concern. The definition of unsafe fish consumption used in this document
(represented by "NONE" in the tables) Is a consumption limit of less than
one meal every 2 months (<6 meals/year) for the monthly tables. Where risk
calculations indicate that six fish mea'ls per year may not be safely eaten, risk
managers may choose to advise the public that no contaminated recreatio'nally
and subsistence caught fish should be consumed from designated waterbodies.
It is the responsibility of risk managers to define safe and unsafe fish consumption
limits, based on a variety of considerations, including adverse health effects and
the consumption patterns of the population of concern. Note: The chemical
contaminant concentrations in fish tissues listed a're only those that fall between
these two limits, and these fish tissue concentrations are not listed in a linear
fashion. Directions and equations for calculating other meal consumption limits not
found in the consumption limit tables are provided in Section 3.
Specific guidance concerning the use of these limits in a fish advisory program is
provided in Volume 3: Overview of Risk Management. Readers are encouraged
to review this and the other documents in the series regarding selection of
4-3
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
consumption limits for their fish advisory programs. Fish advisories involve many
factors that are not represented in these tables, including societal, economic,
nutritional, and cultural impacts. These factors also are discussed in Volume 3.
4.2 Consumption Limit Tables
Table 4-1 provides a directory of the consumption limit tables for all the target
analytes recommended for analysis in fish sampling programs. Table 4-2 sum-
marizes RfD, CSF, and detection limit values used to develop fish consumption
limits Tables 4-3 through 4-64 are consumption limit tables for carcinogenic and
chronic systemic health endpoints for different chemicals. 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 con-
taminant situations and for multiple fish species diets.
4-4
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
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4. RISK-BASED CONSUMPTION LIMIT TABLES
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4-6
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-3. Monthly Consumption Limits for Chronic Systemic Health Endpoints for the
General Population - Arsenic (inorganic)
Chemical
Concentration
in Fish Tissue3
(mg/kg or ppm)
<0.05
0.05
0.06
0.07
0.08
0.09
0.1
0.2
0.3
0.4
0.5
0.6
0.7
0.8
0.9
1
2
3
4
5
6
7
8
9
10
11
>11
Recommended Risk - Based Consumption Limit (meals per month)"
4-oz Meal Size
(0.114 ka)
>30
>30
>30
>30
>30
>30
>30
28
18
14
11
9
8
7
6
5
2
1
1
1
6/yr
6/yr
6/yr
6/yr
6/yr
6/yr
NONE
8-oz Meal Size
(0.227 kq)
>30
>30
>30
>30
>30
>30
28
14
9
7
5
4
4
3
3
2
1
6/yr
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE
12-oz Meal Size
(0.341 kq)
>30
>30
>30
26
23
20
18
9
6
4
3
3
2
2
2
1
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
16-oz Meal Size
(0 454 kq)
>30
28
23
20
17
15
14
7
4
3
2
2
2
1
1
1
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE = No consumption recommended.
6/yr = Consumption of no more than 6 meals per year is recommended.
> 30 = Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 meal per day).
a Instructions for modifying the variables in this table are found in Section 3.3.
Consumption limits are calculated based on an adult body weight of 70 kg and using RfD=3x lO
References for RfDs are found in Section 5.
The detection limit is 5 x 10~3 mg/kg.
All values were rounded down to the nearest whole meal size.
Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD).
When this dose is consumed in less than 1 month (e.g., in a few large meals), the daily dose will
exceed the RfD (see Section 2.3).
4-7
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-4. Monthly Consumption Limits for Chronic Systemic Health Endpoints for
Children - Arsenic (inorganic)
Chemical
Concentration
In Fish Tissue*
(mg/kg or ppm)
<0.02
0.02
0.03
0.04
0.05
0.06
0.07
0.08
0.09
0.1
0.2
0.3
0.4
0.5
0.6
0.7
0.8
0.9
1
2
3
>3
Recommended Risk - Based Consumption Limit (meals per month)"
3-oz Meal Size
(0 085 kq)
>30
>30
>30
>30
>30
25
22
19
17
15
7
5
3
3
2
2
1
1
1
6/yr
6/yr
NONE
4-oz Meal Size
(0.11 4 kq)
>30
>30
>30
29
23
19
16
14
12
11
5
3
2
2
1
1
1
1
1
6/yr
NONE
NONE
8-oz Meal Size
(0.227 kq)
>30
29
19
14
11
9
8
7
6
5
2
1
1
1
6/yr
6/yr
6/yr
6/yr
6/yr .
NONE
NONE
NONE
12-ozMeal Size
(0.341 kq)
>30
19
12
9
7
6
5
4
4
3
1
1
6/yr
6/yr
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE = No consumption recommended.
6/yr = Consumption of no more than 6 meals per year is recommended.
> 30 = Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 meal per day).
* Instructions for modifying the variables in this table are found in Section 3.3.
Consumption limits are calculated based on a body weight of 14.5 kg and using RfD=3x10"4mg/kg/d.
References for Rf Ds are found in Section 5.
The detection limit is 5 x 10"3 mg/kg.
All values were rounded down to the nearest whole meal size.
b Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD).
When this dose is consumed in less than 1 month (e.g., in a few large meals), the daily dose will
exceed the RfD (see Section 2.3).
4-8
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
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-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-6. Monthly Consumption Limits for Chronic Systemic Health Endpoints for the
General Population - Cadmium
Chemical
Concentration
in Fish Tissue8
(mg/kg or ppm)
<0.2
0.2
0.3
0.4
0.5
0.6
0.7
0.8
0.9
1
2
3
4
5
6
7
8
9
10
20
30
>30
Recommended Risk - Based Consumption Limit (meals per month)"
4-oz Meal Size
(0.11 4 kq)
>30
>30
>30
>30
>30
>30
26
23
20
18
9
6
4
3
3
2
2
2
1
6/yr.
6/yr
NONE
8-oz Meal Size
(0.227 kq)
>30
>30
>30
23
18
15
13
11
10
9
4
3
, 2
1
1
1
: 1
. 1
6/yr
NONE
NONE
NONE
12-oz Meal Size
(0.341 kq)
>30
>30
20
15
12
10
8
7
6
6
3
2
1
1
1
6/yr
6/yr
6/yr
6/yr
NONE
NONE
NONE
16-oz Meal Size
(0.454 kq)
23
15
11
9
7
6
5
5
4
2
1
1
6/yr
6/yr
6/yr
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE = No consumption recommended.
6/yr = Consumption of no more than 6 meals per year is recommended.
> 30 = Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 meal per day).
* Instructions for modifying the variables in this table are found in Section 3,3.
Consumption limits are calculated based on an adult body weight of 70 kg and.using RfD=1x1o;3mg/kg/d.
References for RfDs are found in Section 5.
The detection limit is 5 x 10"3 mg/kg.
All values were rounded down to the nearest whole meal size.
b Monthly limits are based on the total dose allowable over a, 1 -month period (based on the RfD).
When this dose is consumed in less than 1 month (e.g., in a few large meals), the daily dose will
exceed the RfD (see Section 2.3).
4-10
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-7. Monthly Consumption Limits for Chronic Systemic Health Endpoints for
Children - Cadmium
Chemical
Concentration
in Fish Tissue3
(mg/kg or ppm)
0.05
0.06
0.07
0.08
0.09
0.1
0.2
0.3
0.4
0.5
0.6
0.7
0.8
0.9
1
2
3
4
5
6
7
8
9
10
>10
Recommended Rjsk - Based Consumption Limit (meals per month)"
3-oz Meal Size
(0.085 ka)
>30
>30
>30
>30
>30
>30
>30
25
17
12
10
8
7
6
5
5
2
1
1
1
6/yr
6/yr
6/yr
6/yr
6/yr
NONE
4-oz Meal Size
(0.11 4 ka)
>30
>30
>30
>30
>30
>30
>30
19
12
9
7
6
5
4
4
3
1
1
6/yr
6/yr
6/yr
6/yr ;
NONE
NONE
NONE
NONE
8-oz Meal Size
(0.227 ka)
>sd
>30
>30
27
24
21
19
9
6
4
3
3
2
2
2
1 •
6/yr
• 6/yr
.NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
12-oz Meal Size
>30
25
21
18
16
14
12
6
4
3
2
2
1
1
1
1
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE = No consumption recommended.
6/yr = Consumption of no more than 6 meals per year is recommended.
> 30 = Although consumption of more than 30 me.als/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 meal per day).
a Instructions for modifying the variables in this table are found in Section 3.3.
Consumption limits are calculated based on a body weight of 14.5 kg and using RfD=1 x1 p"3 mg/kg/d.
References for RfDs are found in Section 5.
The detection limit is 5 x 10"3 mg/kg.
All values were rounded down to the nearest whole meal size. .
"Monthly limits are based pn the total dose allowable over a 1-month period (based on the RfD).
When this dose is consumed in less than 1 month (e.g., in a few large meals), the daily dose will
exceed the RfD (see Section 2.3). .
4-11
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-8. Monthly Consumption Limits for Chronic Systemic Health Endpoints for the
General Population - Methylmercury (IRIS RfD)
Chemical
Concentration
in Fish Tissue*
(mg/kg or ppm)
<0.02
0.02
0.03
0.04
0.05
0.06
0.07
0.08
0.09
0.1
0.2
0.3
0.4
0.5
0.6
0.7
0.8
0.9
1
2
3
Recommended Risk - Based Consumption Limit (meals per month)"
4-oz Meal Size
>30
>30
>30
>30
>30
>30
26
23
20
18
9
6
4
3
3
2
2
2
1
6/yr
6/yr
NONE
8-oz Meal Size
(0.227 ka)
>30
>30
>30
23
18
15
13
11
10
9
4
3
2
1
1
1
1
1
6/vr
NONE
NONE
NONE
12-oz Meal Size
(0.341 ka)
>30
20
15
12
10
8
7
6
6
3
2
1
. 1
1
6/yr
6/yr
6/yr
6/yr
NONE
NONE
NONE
16-ozMeal Size
(0.454 ka)
23
15
11
9
7
6
5
5
4
2
1
1
6/yr
6/yr
6/yr
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE = No consumption recommended.
6/yr = Consumption of no more than 6 meals per year is recommended.
> 30 = Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 meal per day).
* Instructions for modifying the variables in this table are found in Section 3.3.
Consumption limits are calculated based on an adult body weight of 70 kg and using Rf D=1 x10 mg/kg/d.
References for RfDs are found in Section 5.
The detection limit is 1 x 10"3 mg/kg.
AH values were rounded down to the nearest whole meal size.
b Monthly limits are based on the total dose allowable over a 1 -month period (based on the RfD).
When this dose is consumed in less than 1 month (e.g., in a few large meals), the daily dose will
exceed the RfD (see Section 2.3).
4-12
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-9. Monthly Consumption Limits for Developmental Health Endpoints for Women of
Reproductive Age - Methylmercury (IRIS RfD)
Chemical
Concentration
in Fish Tissue9
(mg/kg or ppm)
<0.02
0.02
0.03
0.04
0.05
0.06
0.07
0.08
0.09
0.1
0.2
0.3
0.4
0.5
0.6
0.7
0.8
0.9
1
2
3
>3
Recommended Risk - Based Consumption Limit (meals per month)"
4-oz Meal Size
(0.114kq)
>30
>30
>30
>30
>30
>30
26
23
20
18
9
6
4
3
3
2
2
2
1
6/yr
6/yr
NONE
8-oz Meal Size
(0.227 kq)
>30
>30
>30
23
18
15
13
11
10
9
4
3
2
1
1
1
1
1
6/yr
NONE
NONE
NONE
12-oz Meal Size
(0.341 kq)
>30
>30
20
15
12
10
8
7
6
6
3
2
1
1
1
6/yr
6/yr
6/yr
6/yr
NONE
NONE
NONE
16-oz Meal Size
(0 454 kq)
>30
23
15
11
9
7
6
5
5
4
2
1
1
6/yr
6/yr
6/yr
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE = No consumption recommended.
6/yr = Consumption of no more than 6 meals per year is recommended.
> 30 = Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 meal per day).
a Instructions for modifying the variables in this table are found in Section 3.3.
Consumption limits are calculated based on an adult body weight of 70 kg and using RfD=1x10r4mg/kg/d.
References for RfDs are found in Section 5.
The detection limit is 1 x 10~3 mg/kg.
All values were rounded down to the nearest whole meal size. .
b Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD).
When this dose is consumed in less than 1 month (e.g., in a few large meals), the daily dose will
exceed the RfD (see Section 2.3).
4-13
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-10. Monthly Consumption Limits for Developmental Health Endpoints for Children -
Methylmercury (IRIS RfD)
Chemical
Concentration
in Fish Tissue8
(mg/kg or ppm)
<0.005
0.005
0.006
0.007
0.008
0.009
0.01
0.02
0.03
0.04
0.05
0.06
0.07
0.08
0.09
0.1
0.2
0.3
0.4
0.5
0.6
0.7
0.8
0.9
1
>-j
Recommended Risk - Based Consumption Limit (meals per month)"
3-oz Meal Size
(0 085 kq)
>30
>30
>30
>30
>30
>30
>30
25
17
12
10
8
7
6
5
5
2
1
1
1
6/yr
6/vr
6/vr
6/yr
6/yr
NONE
4-oz Meal Size
(0.114kg)
>30
>30
>30
>30
>30
>30
>30
19
12
9
7
6
5
4
4
3
1
1
6/yr
6/yr
6/yr
6/yr
NONE
NONE
NONE
NONE
8-oz Meal Size
(0.227 kg)
>30
>30
>30
27
24
21
19
9
6
4
3
3
2
2
2
1
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
12-ozMeal Size
(0.341 kg)
>30
25
21
18
16
14
12
6
4
3
2
2
1
1
1
1
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE = No consumption recommended.
6/yr = Consumption of no more than 6 meals per year is recommended.
> 30 = Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 meal per day).
* Instructions for modifying the variables in this table are found in Section 3.3.
Consumption limits are calculated based on a body weight of 14.5 kg and using RfD=1x10"4mg/kg/d.
References for RfDs are found in Section 5.
The detection limit is 1 x 10~3 mg/kg.
All values were rounded down to the nearest whole meal size.
b Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD).
When this dose is consumed in less than 1 month (e.g., in a few large meals), the daily dose will
exceed the RfD (see Section 2.3).
4-14
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-11. Monthly Consumption Limits for Chronic Systemic Health Endpoints for the
General Population - Selenium
Chemical
Concentration
in Fish Tissue8
(mg/kg or ppm)
<0.8
0.8
0.9
1
2
3
4
5
6
7
8
9
10
20
30
40
50
60
70
80
90
100
120
140
160
180
>180
Recommended Risk - Based Consumption Limit (meals per month)"
4-oz Meal Size
(0.11 4 kq)
>30
>30
>30
>30
>30
>30
23
18
15
13
11
10
9
4
3
2
1
1
1
1
1
6/yr
6/yr
6/yr
6/yr
6/yr
NONE
8-oz Meal Size
(0.227 ka)
>30
>30
>30
>30
23
15
11
' 9
7
6
5
5
4
2
1
1
6/yr
6/yr
6/yr
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
12-ozMeal Size
(0.341 kq)
>30
, >30
, >30 •
>30 '
15
10
7
6
5
4
3
3
3
1
1
6/yr
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
16-oz Meal Size
(0.454 kg)
>30
;- • 29
26
23
11 .
7
5
4
3
3
2
2
2
1
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE = No consumption recommended.
6/yr == Consumption of no more than 6 meals per year is recommended.
> 30 = Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 meal per day).
a Instructions for modifying the variables in this table are found in Section 3.3.
Consumption limits are calculated based on an adult body weight of 70 kg and using RfD=5x10"3 mg/kg/d.
References for RfDs are found in Section 5.
The detection limit is 2 x 10"2 mg/kg.
All values were rounded down to the nearest whole meal size.
b Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD).
When this dose is consumed in less than 1 month (e.g., in a few large meals), the daily dose will
exceed the RfD (see Section 2.3).
4-15
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-12. Monthly Consumption Limits for Chronic Systemic Health Endpoints for
Children - Selenium
Chemical
Concentration
in Fish Tissue*
(mg/kg or ppm)
<0.3
0.3
0.4
0.5
0.6
0.7
0.8
0.9
1
2
3
4
5
6
7
8
9
10
20
30
40
50
Recommended Risk - Based Consumption Limit (meals per month)"
3-oz Meal Size
(0.085 kq)
>30
>30
>30
>30
>30
>30
>30
28
25
12
8
6
5
4
3
3
2
2
1
6/yr
6/vr
6/yr
NONE
4-oz Meal Size
(0.11 4 kq)
>30
>30
>30
>30
>30
27
24
21
19
9
6
4
3
3
2
2
2
1
6/vr
6/yr
NONE
NONE
NONE
8-oz Meal Size
(0.227 ka)
>30
>30
24
19
16
13
-12
10
9
4
3
2
1
1
1
1
1
6/yr
NONE
NONE
NONE
NONE
NONE
12-oz Meal Size
(0.341 ka)
21
16
12
10
9
8
7
6
3
2
1
1
1
6/yr
6/yr
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE = No consumption recommended.
6/yr a* Consumption of no more than 6 meals per year is recommended.
> 30 = Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 meal per day).
* Instructions for modifying the variables in this table are found in Section 3.3.
Consumption limits are calculated based on a body weight of 14.5 kg and using RfD=5x10"3 mg/kg/d.
References for RfDs are found in Section 5.
The detection limit is 2 x 10"2 mg/kg.
All values were rounded down to the nearest whole meal size.
b Monthly limits are based on the total dose allowable over a 1 -month period (based on the RfD).
When this dose is consumed in less than 1 month (e.g., in a few large meals), the daily dose will
exceed the RfD (see Section 2.3).
4-16
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-13. Monthly Consumption Limits for Chronic Systemic Health Endpoints for the
General Population - Tributyltin Oxide
Chemical
Concentration
in Fish Tissue9
(mg/kg or ppm)
<0.004
0.005
0.006
0.007
0.008
0.009
0.01
0.02
0.03
0.04
0.05
0.06
0.07
0.08
0.09
0.1
0.2
0.3
0.4
0.5
0.6
0.7
0.8
0.9
1
>1
. Recommended Risk - Based Consumption Limit (meals per month)b
4-oz Meal Size
(0.11 4 kq)
>30
>30
>30
>30
>30
>30
>30
28
18
14
11
9
8
7
6
5
2
1
1
1
6/yr
6/yr
6/yr
6/yr
6/yr
NONE
8-oz Meal Size
(0.227 kq)
>30
>30
>30
>30
>30
>30
28
14
9
7
5
4
4
3
3
2
1
6/yr
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
12-oz Meal Size
(0.341 kq)
>30
>30
>30
26
23
20
18
9
6
4
3
3
2
2
2
1
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
16-oz Meal Size
(0.454 kq)
>30
28
23
20
17
15
14
7
4
3
2
2
2
1
1
1
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE = No consumption recommended.
6/yr = Consumption of no more than 6 meals per year is recommended.
> 30 = Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 meal per day).
a Instructions for modifying the variables in this table are found in Section 3.3.
Consumption limits are calculated based on an adult body weight of 70 kg and using RfD=3x10"5 mg/kg/d.
References for RfDs are found in Section 5.
The detection limit is 1 x 10 "3 mg/kg.
All values were rounded down to the nearest whole meal size.
b Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD).
When this dose is consumed in less than 1 month (e.g., in a few large meals), the daily dose will
exceed the RfD (see Section 2.3).
4-17
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-14. Monthly Consumption Limits for Chronic Systemic Health Endpoints for
Children - Tributyltin Oxide
Chemical
Concentration
in Fish Tissue"
(mg/kg or ppm)
<0.002
0.002
0.003
0.004
0.005
0.006
0.007
0.008
0.009
0.01
0.02
0.03
0.04
0.05
0.06
0.07
0.08
0.09
0.1
0.2
0.3
Recommended Risk - Based Consumption Limit (meals per month)
3-oz Meal Size
(0.085 kq)
>30
>30
>3Q
>30
>30
25
22
19
17
15
7
5
3
3
2
2
1
1
1
6/vr
6/yr
NONE
4-oz Meal Size
(0.11 4 kq)
>30
>30
>30
29
23
19
16
14
12
11
5
3
2
2
1
1
1
1
1
6/yr
NONE
NONE
8-oz Meal Size
(0.227 kq)
>30
29
19
14
11
9
8
7
6
5
2
1
1
1
6/yr
6/yr
6/yr
6/yr
6/yr
NONE
NONE
NONE
12-ozMeal Size
(0.341 kq)
19
12
9
7
6
5
4
4
3
1
1
6/yr
6/yr
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE = No consumption recommended.
6/yr = Consumption of no more than 6 meals per year is recommended.
> 30 = Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 meal per day).
* Instructions for modifying the variables in this table are found in Section 3.3.
Consumption limits are calculated based on a body weight of 14.5 kg and using Rf D=3x10"5 mg/kg/d.
References for RfDs are found in Section 5.
The detection limit is 1 x 10"3 mg/kg.
All values were rounded down to the nearest whole meal size.
b Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD).
When this dose is consumed in less than 1 month (e.g., in a few large meals), the daily dose will
exceed the RfD (see Section 2.3).
4-18
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-15. Monthly,Consumption Limits for Chronic Systemic Health Endpoints for the
General Population - Chlordane
Chemical
Concentration
in Fish Tissue8
(mg/kg or ppm)
<0.01
0.01
0.02
0.03
0.04
0.05
0.06
0.07
0.08
0.09
0.1
0.2
0.3
0.4
0.5
0.6
0.7
0.8
0.9
1
2
>2
Recommended Bisk - Based Consumption Limit (meals per month)"
4-oz Meal Size
(0.11 4 kq)
>30
>30
>30
>30
28
22
18
16
1,4
12
11
5
3
2
2
1
1
1
1
1
6/yr
NONE
8-oz Meal Size I 12-oz Meal Size I 16-oz Meal Size
(0.227 kq) | (0 341 kq) I (0 454 ka\ '
' ">30
>30
28
18
14
11
9
8
7
6
5
2
1
1
1
6/yr
6/yr
6/yr
6/vr
6/yr
NONE
NONE
>30 ' "'•
>30
18
12
9
7
6
5
4
4
3
1
1
6/vr
6/vr
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
>"30 '•'
28
14
9
7
5
4
4
3
3
"2
1 :
6/vr
6/vr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE =No consumption recommended.
;6/yr = Consumption of no more than 6 meals per year is recommended.
> 30 = Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption tp 30 meals in 1 month (1 meal-per day).
a Instructions for modifying the variables in this table are found in Section 3.3.
Consumption limits are calculated based on an. adult body weight of 70 kg and using RfD=6x10"5mg/kg/d.
References for RfDs are found in Section 5.
The detection limit is 1 x 10"3 mg/kg.
All values were rounded down to the nearest whole meal size.
Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD).
When this dose is consumed in less than 1 month (e.g., in a few large meals), the daily dose will
exceed the RfD,(see Section 2.3).
4-19
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-16. Monthly Consumption Limits for Chronic Systemic Health Endpoints for
Children - Chlordane
Chemical
Concentration
in Fish Tissue*
(mg/kg or ppm)
<0.003
0.003
0.004
0.005
0.006
0.007
0.008
0.009
0.01
0.02
0.03
0.04
0.05
0.06
0.07
0.08
0.09
0.1
0.2
0.3
0.4
0.5
0.6
Recommended Risk - Based Consumption Limit (meals per month)"
3-oz Meal Size
>30
>30
>30
>30
>30
>30
>30
>30
>30
15
10
7
6
5
4
3
3
3
1
1
6/vr
6/vr
6/yr
4-oz Meal Size
(0.11 4 kq)
>30
>30
>30
>30
>30
>30
29
25
23
11
7
5
4
3
3
2
2
2
1
6/yr
6/vr
NONE
NONE
NONE
8-oz Meal Size
(0.227 ka)
>30
29
23
19
16
14
12
11
5
3
2
2
1
1
1
1
1
6/yr
NONE
NONE
NONE
NONE
NONE
12-oz Meal Size
(0.341 kcrt
25
19
15
12
11
9
8
7
3
2
1
1
1
1
6/yr
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE s No consumption recommended.
6/yr = Consumption of no more than 6 meals per year is recommended.
> 30 ~ Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 meal per day).
* Instructions for modifying the variables in this table are found in Section 3.3. ^
Consumption limits are calculated based on a body weight of 14.5 kg and using RfD=6x10" mg/kg/d.
References for RfDs are found in Section 5.
The detection limit is 1 x 10'3 mg/kg.
All values were rounded down to the nearest whole meal size.
b Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD).
When this dose is consumed in less than 1 month (e.g., in a few large meals), the daily dose will
exceed the RfD (see Section 2.3).
4-20
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
'5
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4-21
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-18. Monthly Consumption Limits for Chronic Systemic Health Endpoints for the
General Population - DDT
Chemical
Concentration
in Fish Tissue8
(mg/kg or ppm)
<0.08
0.08
0.09
0.1
0.2
0.3
0.4
0.5
0.6
0.7
0.8
0.9
1
2
3
4
5
6
7
8
9
10
12
14
16
18
Recommended Risk - Based Consumption Limit (meals per month)"
4-oz Meal Size
>30
>30
>30
>30
>30
>30
23
18
15
13
11
10
g
4
3
2
1
1
1
1
1
6/vr
6/yr
6/yr
6/vr
6/vr
8-oz Meal Size
(0.227 kq)
>30
>30
>30
>30
23
15
11
9
7
6
5
5
4
2
1
1
6/yr
6/yr
6/yr
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
12-oz Meal Size
(0.341 kq)
>30
>30
>30
15
10
7
6
5
4
3
3
3
1
1
6/yr
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
16-oz Meal Size
(0.454 kq)
29
26
23
11
7
5
4
3
3
2
2
2
1
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE = No consumption recommended.
6/yr = Consumption of no more than 6 meals per year is recommended.
> 30 = Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 meal per day).
a Instructions for modifying the variables in this table are found in Section 3.3.
Consumption limits are calculated based on an adult body weight of 70 kg and using RfD=5x10'4 mg/kg/d.
References for RfDs are found in Section 5.
The detection limit is 1 x 10"4 mg/kg.
All values were rounded down to the nearest whole meal size.
"Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD).
When this dose is consumed in less than 1 month (e.g., in a few large meals), the daily dose will
exceed the RfD (see Section 2.3).
4-22
-------
4. RISK-BASED CONSUMPTION LfMIT TABLES
Table 4-19. Monthly Consumption Limits for Chronic Systemic Health Endpoints for
Children - DDT
Chemical
Concentration
in Fish Tissue8
(mg/kg or ppm)
<0.03
0.03
0.04
0.05
0.06
0.07
0.08
0.09
0.1
0.2
0.3
0.4
0.5
0.6
0.7
0.8
0.9
1
2
3
4
5
>5
Recommended Risk - Based Consumption Limit (meals per month)"
3-oz Meal Size
(0.085 kq)
>30
>30
>30
>30
>30
>30
>30
28
25
12
8
6
5
4
3
3
2
2
1
6/yr
6/yr
6/yr
NONE
4-oz Meal Size
(0.114kg)
>30
>30
>30
>30
>30
27
24
21
19
9
6
4
3
3
2
2
2
1
6/yr
6/yr
NONE
NONE
NONE
8-oz Meal Size
(0.227 kq)
>30
>30
24
19
16
13
12
10
9
4
3
2
1
1
1
1
1
6/yr
NONE
NONE
NONE
NONE
NONE
12-oz Meal Size
(0 341 kq)
>30
21
16
12
10
9
8
7
6
3
2
1
1
1
6/yr
6/yr
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE = No consumption recommended.
6/yr = Consumption of no more than 6 meals per year is recommended.
> 30 = Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 meal per day).
a Instructions for modifying the variables in this table are found in Section 3.3.
Consumption limits are calculated based on a body weight of 14.5 kg and using RfD=5x O^mg/kg/d.
References for RfDs are found in Section 5.
The detection limit is 1 x 10~4 mg/kg.
All values were rounded down to the nearest whole meal size.
b Monthly limits are based on the total dose allowable over a 1 -month period (based on the RfD).
When this dose is consumed in less than 1 month (e.g., in a few large meals), the daily dose will
exceed the RfD (see Section 2.3).
4-23
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
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4-24
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-21. Monthly Consumption Limits for Chronic Systemic Health Endpoints for the
General Population - Dicofol
Chemical
Concentration
in Fish Tissue3
(mg/kg or ppm)
<0.2
0.2
0.3
0.4
0.5
0.6
0.7
0.8
0.9
1
2
3
4
5
6
7
8
9
10
15
20
25
30
35
>35
Recommended Risk - Based Consumption Limit (meals per month)b
4-oz Meal Size
(0.11 4 kq)
>30
>30
>30
>30
>30
>30
26
23
20
18
9
6
4
3
3
2
2
2
1
1
6/yr
6/yr
6/yr
6/yr
NONE
8-oz Meal Size
(0.227 kq)
>30
>30
>30
23
18
15
13
11
10
9
4
3
2
1
1
1
1
1
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
12-oz Meal Size
(0.341 kq)
>30
>30
20
15
12
10
8
7
6
6
3
2
1
1
1
6/yr
6/yr
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
16-oz Meal Size
(0 454 kq)
>30
23
15
11
9
7
6
5
5
4
2
1
1
6/yr
6/yr
6/yr
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE .
NONE .
NONE
NONE = No consumption recommended.
6/yr = Consumption of no more than 6 meals per year is recommended.
> 30 = Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 meal per day).
a Instructions for modifying the variables in this table are found in Section 3.3.
Consumption limits are calculated based on an adult body weight of 70 kg and using RfD=1x10~3mg/kg/d.
References for RfDs are found in Section 5.
The detection limit is 1 x 10"2 mg/kg.
All values were rounded down to the nearest whole meal size.
b Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD).
When this dose is consumed in less than 1 month (e.g., in a few large meals), the daily dose will
exceed the RfD (see Section 2.3).
4-25
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-22. Monthly Consumption Limits for Chronic Systemic Health Endpoints for
Children - Dicofol
Chemical
Concentration
in Fish Tissue*
(mg/kg or ppm)
<0.05
0.05
0.06
0.07
0.08
0.09
0.1
0.2
0.3
0.4
0.5
0.6
0.7
0.8
0.9
1
2
3
4
5
6
7
8
9
10
Recommended Risk - Based Consumption Limit (meals per month)"
3-oz Meal Size
(0.085 kq)
>30
>30
>30
>30
>30
>30
>30
25
17
12
10
8
7
6
' 5
5
2
1
1
1
•6/vr
6/vr
6/vr
6/vr
6/yr
NONE
4-oz Meal Size
(0.11 4 kd)
>30
>30
>30
>30
>30
>30
>30
19
12
9
7
6
5
4
4
3
1
1
6/yr
6/yr
6/yr
6/yr
NONE
NONE
NONE
NONE
8-oz Meal Size
(0.227 kq)
>30,
>30
27
24
21
, 19
. 9
6
4
3
3
2
2
2
1
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
12-oz Meal Size
(0.341 kq)
25
21
18
16
14 ,
12
6
4
3
2
2
1
1
1
1
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE .
NONE = No consumption recommended.
6/yr = Consumption of no more than 6 meals per year is recommended.
> 30 = Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 meal per day).
* Instructions for modifying the variables in this table are found in Section 3.3.
Consumption limits are calculated based on a body weight of 14.5 kg and using RfD=1x10'3 mg/kg/d.
References for RfDs are found in Section 5.
The detection limit is 1 x 10"2 mg/kg.
All values were rounded down to the nearest whole meal size.
b Monthly limits are based on the total dose allowable over a 1 -month period (based on the RfD).
When this dose is consumed in less than 1 month (e.g., in a few large meals), the daily dose will
exceed the RfD (see Section 2.3).
4-26
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
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in Fish Tissue3
(mg/kg or ppm)
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4-27
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-24. Monthly Consumption Limits for Chronic Systemic Health Endpoints for the
General Population - Dieldrin
Chemical
Concentration
In Fish Tissue8
(mg/kg or ppm)
<0.008
0.008
0.009
0.01
0.02
0.03
0.04
0.05
0.06
0.07
0.08
0.09
0.1
0.2
0.3
0.4
0.5
0.6
0.7
0.8
0.9
1
Recommended Risk - Based Consumption Limit (meals per month)"
4-oz Meal Size
>30
>30
>30
>30
>30
>30
23
18
15
13
11
10
9
4
3
2
1
1
1
1
1
6/vr
8-oz Meal Size
(0.227 kq)
>30
>30
>30
>30
23
15
11
9
7
6
5
5
4
2
1
1
6/yr
6/yr
6/yr
6/yr
6/yr
NONE
NONE
12-oz Meal Size
(0.341 kq)
>30
>30
>30
15
10
7
6
5
4
3
3
3
1
1
6/yr
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
16-oz Meal Size
(0.454 ka)
29
26
23
11
7
5
4
3
3
2
2
2
1
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE = No consumption recommended.
6/yr = Consumption of no more than 6 meals per year is recommended.
> 30 = Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 meal per day).
* Instructions for modifying the variables in this table are found in Section 3.3. ^
Consumption limits are calculated based on an adult body weight of 70 kg and using RfD=5x10" mg/kg/d.
References for RfDs are found in Section 5.
The detection limit is 1 x 10"4 mg/kg.
All values were rounded down to the nearest whole meal size.
b Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD).
When this dose is consumed in less than 1 month (e.g., in a few large meals), the daily dose will
exceed the RfD (see Section 2.3).
4-28
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-25. Monthly Consumption Limits for Chronic Systemic Health Endpoints for
Children - Dieldrin
Concentration
in Fish Tissue8
(mg/kg or ppm)
<0.003
0.003
0.004
0.005
0.006
0.007
0.008
0.009
0.01
0.02
0.03
0.04
0.05
0.06
0.07
0.08
0.09
0.1
0.2
0.3
0.4
, 0.5
>0.5
Recommended Risk - Based Consumption Limit (meals per month)"
3-oz Meal Size
(0.085 kq)
>30
>30
>30
>30
>30
>30
>30
28
25
12
8
6
5
4
3
3
2
2
1
6/yr
6/yr
6/yr
NONE
4-oz Meal Size
(0.11 4 kq)
>30
>30
>30
>30
>30
27
24
21
19
9
6
4
3
3
2
2
2
1
6/yr
6/yr
NONE
NONE
NONE
8-oz Meal Size I 12-oz Meal Size
(0 227 kq) I (0 341 ka)
>30
>30
24
19
16
13
12
10
9
4
3
2
1
1
1
1
1
6/yr
NONE
NONE
NONE
NONE
NONE
>30
21
16
12
10
9
8
•7
6
3
2
1
1
1
6/yr
6/vr
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE = No consumption recommended.
6/yr = Consumption of no more than 6 meals per year is recommended.
> 30 = Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 meal per day).
a Instructions for modifying the variables in this table are found in Section 3.3.
Consumption limits are calculated based on a body weight of 14.5 kg and using RfD=5x10"5 mg/kg/d.
References for RfDs are found in Section 5.
The detection limit is 1 x 10"4 mg/kg.
All values were rounded down to the nearest whole meal size.
Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD).
When this dose is consumed in less than 1 month (e.g., in a few large meals), the daily dose will
exceed the RfD (see Section 2.3).
4-29
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-26. Monthly Consumption Limits for Carcinogenic Health Endpoints
for the General Population - Dieldrin
Recommended Risk - Based Consumption Limit (meals per month1
5
t
9
0
5
E
D
1
3
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8-oz Meal Size (0.227 kg) |
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4-30
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-27. Monthly Consumption Limits for Chronic Systemic Health Endpoints for the
General Population - Endosulfan
Chemical
Concentration
in Fish Tissue8
(mg/kg or ppm)
1
2
3
4
5
6
7
8
9
10
20
30
40
50
60
70
80
90
100
120
140
160
180
200
220
>220
Recommended Risk - Based Consumption Limit (meals per month)b
4-oz Meal Size
(0.11 4 kcrt
>30
>30
>30
28
22
18
16
14
12
11
5
3
2
2
6/yr
6/yr
6/yr
6/yr
6/yr
6/yr
NONE
8-oz Meal Size
(0.227 ka)
>30
>30
28
18
14
11
9
8
7
6
5
2
1
1
1
6/yr
6/yr
6/yr
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE
12-oz Meal Size
(0.341 ka)
>30
>30
18
12
9
7
6
5
4
4
3
1
1
6/yr
6/yr
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
16-oz Meal Size
>30
28
14
9
7
5
4
4
3
3
2
1
6/yr
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE = No consumption recommended.
• 6/yr = Consumption of no more than 6 meals per year is recommended.
> 30 = Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 meal per day).
a Instructions for modifying the variables in this table are found in Section 3.3.
Consumption limits are calculated based on an adult body weight of 70 kg and using RfD=6x10"3 mg/kg/d.
References for RfDs are found in Section 5.
The detection limit is 5 x 10~3 mg/kg.
All values were rounded down to the nearest whole meal size.
Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD).
When this dose is consumed in less than 1 month (e.g., in a few large meals), the daily dose will
exceed the RfD (see Section 2.3).
4-31
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-28. Monthly Consumption Limits for Chronic Systemic Health Endpoints for
Children - Endosulfan
Chemical
Concentration
In Fish Tissue8
(mg/kg or ppm)
<0.3
0.3
0.4
0.5
0.6
0.7
0.8
0.9
1
2
3
4
5
6
7
8
g
10
20
30
40
50
60
Recommended Risk - Based Consumption Limit (meals per month)"
3-oz Meal Size
>30
>30
>30
>30
>30
>30
>30
>30
>30
15
10
7
6
5
4
3
3
3
1
1
6/yr
6/yr
6/vr
4-oz Meal Size
(0.114kq)
>30
>30
>30
>30
>30
>30
29
25
23
11
7
5
4
3
3
2
2
2
1
6/yr
6/yr
NONE
NONE
NONE
8-oz Meal Size
(0.227 kq)
>30
29
23
19
16
14
12
11
5
3
2
2
6/yr
NONE
NONE
NONE
NONE
NONE
12-ozMeal Size
(0.341 ka)
25
19
15
12
11
9
8
7
3
2
1
1
1
1
6/yr
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE = No consumption recommended.
6/yr = Consumption of no more than 6 meals per year is recommended.
> 30 = Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 meal per day).
* Instructions for modifying the variables in this table are found in Section 3.3. ^
Consumption limits are calculated based on a body weight of 14.5 kg and using RfD=6x10" mg/kg/d.
References for RfDs are found in Section 5.
The detection limit is 5 x 10"3 mg/kg.
All values were rounded down to the nearest whole meal size.
* Monthly limits are based on the total dose allowable over a 1 -month period (based on the Rf D).
When this dose is consumed in less than 1 month (e.g., in a few large meals), the daily dose will
exceed the RfD (see Section 2.3).
4-32
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-29. Monthly Consumption Limits for Chronic Systemic Health Endpoints for the
General Population - Endrin
Concentration
in Fish Tissue8
(mg/kg or ppm)
0.05
0.06
0.07
0.08
0.09
0.1
0.2
0.3
0.4
0.5
0.6
0.7
0.8
0.9
1
2
3
4
5
6
7
8
9
10
11
>11
Recommended Risk - Based Consumption Limit (meals per month)"
4-oz Meal Size
(0.11 4 kert
>30
>30
>30
>30
>30
>30
28
18
14
11
9
8
7
6
5
2
1
1
1
6/yr
6/yr
6/yr
6/yr
6/yr
6/yr
NONE
8-oz Meal Size
(0.227 kcrt
>30
>30
>30
>30
>30
>30
28
14
9
7
5
4
4
3
3
2
1
6/yr
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE
12-oz Meal Size
>30
>30
>30
26
23
20
18
9
6
4
3
3
2
2
2
1
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
16-oz Meal Size
>30
28
23
20
17
15
14
7
4
3
o
2
2
1
1
1
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE = No consumption recommended.
6/yr = Consumption of no more than 6 meals per year is recommended.
> 30 = Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 meal per day).
a Instructions for modifying the variables in this table are found in Section 3.3.
Consumption limits are calculated based on an adult body weight of 70 kg and using RfD=3x10'4 ma/ka/d
References for RfDs are found in Section 5. '
The detection limit is 1 x 10~3 mg/kg.
All values were rounded down to the nearest whole meal size.
* Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD)
When this dose is consumed in less than 1 month (e.g., in a few large meals), the daily dose will
exceed the RfD (see Section 2.3).
4-33
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-30. Monthly Consumption Limits for Chronic Systemic Health Endpoints for
Children - Endrin
Chemical
Concentration
in Fish Tissue"
(mg/kg or ppm)
<0.02
0.02
0.03
0.04
0.05
0.06
0.07
0.08
0.09
0.1
0.2
0.3
0.4
0.5
0.6
0.7
0.8
0.9
1
2
3
Recommended Risk - Based Consumption Limit (meals per month)"
3-oz Meal Size
(0.085 ka)
>30
>30
>30
>30
>30
25
22
19
17
15
7
5
3
3
2
2
1
1
1
6/yr
6/yr
4-oz Meal Size
(0.11 4 ka)
>30
>30
>30
29
23
19
16
14
12
11
5
3
2
2
6/yr
NONE
NONE
8-oz Meal Size
(0.227 ka)
>30
29
19
14
11
9
8
7
6
5
2
1 '
1
1
6/yr
6/yr
6/yr
6/yr
6/yr
NONE
NONE
NONE
12-oz Meal Size
(0.341 ka)
>30
19
12
9
7
6
5
4
4
3
1
1
6/yr
6/yr
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE = No consumption recommended.
6/yr = Consumption of no more than 6 meals per year is recommended.
> 30 = Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 meal per day).
" Instructions for modifying the variables in this table are found in Section 3.3. ^
Consumption limits are calculated based on a body weight of 14.5 kg and using RfD=3x1Q- mg/kg/d.
References for RfDs are found in Section 5.
The detection limit is 1 x 10~3 mg/kg.
All values were rounded down to the nearest whole meal size.
b Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD).
When this dose is consumed in less than 1 month (e.g., in a few large meals), the daily dose will
exceed the RfD (see Section 2.3).
4-34
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-31. Monthly Consumption Limits for Chronic Systemic Health Endpoints for the
General Population - Heptachlor Epoxide
Concentration
in Fish Tissue3
(mg/kg or ppm)
<0.002
0.002
0.003
0.004
0.005
0.006
0.007
0.008
0.009
0.01
0.02
0.03
0.04
0.05
0.06
0.07
0.08
0.09
.0.1
0.2
0.3
0.4
>0.4
Recommended Risk - Based Consumption Limit (meals per month)"
4-oz Meal Size
(0.114 kg)
>30
>30
>30
>30
>30
>30
>30
30
26
24
12
8
6
4
4
3
3
2
2
1
6/yr
6/yr
NONE
8-oz Meal Size
(0.227 ka)
>30
>30
>30
30
24
20
17
15
. 13
12
6
4
3
2
2
1
1
1
1
6/yr
NONE
NONE
NONE
12-oz Meal Size
(0.341 kg)
>30
>30
27
20
16
13
11
10
9
8
4
2
2
1
1
1
1
6/yr
6/yr
NONE
NONE
NONE
NONE
=?^^=. ,
16-oz Meal Size
(0.454ka)
>30
30
20
15
12
10
8
7
6
6
3
2
1
1
1
6/yr
6/yr
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE = No consumption recommended.
6/yr = Consumption of no more than 6 meals per year is recommended.
> 30 = Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 mealper day). >
a Instructions for modifying the variables in this table are found in Section 3.3.
Consumption limits are calculated based on an adult body weight of 70 kg and using RfD=1 3x10"5 mo/ka/d
, References for RfDs are found in Section 5. ' '
The detection limit is 1 x 10"4 mg/kg.
All values were rounded down to the nearest whole meal size. '
Monthly limits are based on the total dose allowable over a 1 -month period (based on the RfD).
When this dose is consumed in less than 1 month (e.g., in a few large meals), the daily dose will
exceed the RfD (see Section 2.3).
4-35
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-32. Monthly Consumption Limits for Chronic Systemic Health Endpoints for
Children - Heptachlor Epoxide
Chemical
Concentration
in Fish Tissue*
(mg/kg or ppm)
<0.0006
0.0006
0.0007
0.0008
0.0009
0.001
0.002
0.003
0.004
0.005
0.006
0.007
0.008
0.009
0.01
0.02
0.03
0.04
0.05
0.06
0.07
0.08
0.09
0.1
Recommended Risk - Based Consumption Limit (meals per month)"
3-oz Meal Size
>30
>30
>30
>30
>30
>30
>30
22
16
13
11
g
8
7
6
3
2
1
1
1
6/yr
6/yr
6/yr
6/vr
4-oz Meal Size
(0.11 4 kq)
>30
>30
>30
>30
>30
25
16
•12
10
8
7
6
5
5
2
1
1
1
6/yr
6/yr
6/yr
6/yr
6/yr
NONE
8-oz Meal Size
(0.227 kq)
>30
>30
>30
28
25
12
8
6
5
4
3
3
2
2
1
6/yr '
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
12-oz Meal Size
(0.341 kq)
28
24
21
18
16
8
5
4
3
2
2
2
1
1
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE = No consumption recommended.
6/yr = Consumption of no more than 6 meals per year is recommended.
> 30 = Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 meal per day).
" Instructions for modifying the variables in this table are found in Section 3.3. __
Consumption limits are calculated based on a body weight of 14.5 kg and using RfD=1.3x10" mg/kg/d.
References for RfDs are found in Section 5.
The detection limit is 1 x 10 ^ mg/kg.
All values were rounded down to the nearest whole meal size.
b Monthly limits are based on the total dose allowable over a 1 -month period (based on the RfD).
When this dose is consumed in less than 1 month (e.g., in a few large meals), the daily dose will
exceed the RfD (see Section 2.3).
4-36
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
1
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er year is
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4-37
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-34. Monthly Consumption Limits for Chronic Systemic Health Endpoints for the
General Population - Hexachlorobenzene
Chemical
Concentration
in Fish Tissue3
(mg/kg or ppm)
<0.2
0.2
0.3
0.4
0.5
0.6
0.7
0.8
0.9
1
2
3
4
5
6
7
8
9
10
15
20
25
Recommended Risk - Based Consumption Limit (meals per month)
4-oz Meal Size
>30
>30
>30
>30
29
24
21
18
16
14
7
4
3
2
2
2
1
1
1
6/yr
6/yr
6/yr
8-oz Meal Size
(0.227 kq)
>30
>30
25
18
15
12
10
9
8
7
3
2
1
1
1
1
6/yr
6/yr
6/yr
6/yr
NONE
NONE
NONE
12-oz Meal Size
(0.341 kq)
24
16
12
9
8
7
6
5
4
2
1
1
6/yr
6/yr
6/yr
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
16-oz Meal Size
(0.454 kq)
18
12
9
7
6
5
4
4
3
1
1 '
6/yr
6/yr
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE = No consumption recommended.
6/yr = Consumption of no more than 6 meals per year is recommended.
> 30 = Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 meal per day).
* Instructions for modifying the variables in this table are found in Section 3.3. ^
Consumption limits are calculated based on an adult body weight of 70 kg and using RfD=8x1Q- mg/kg/d.
References for RfDs are found in Section 5.
The detection limit is 1 x 10"4 mg/kg.
All values were rounded down to the nearest whole meal size.
b Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD).
When this dose is consumed in less than 1 month (e.g., in a few large meals), the daily dose will
exceed the RfD (see Section 2.3).
4-38
-------
4. RISK-BASED CONSUMPTION UMIT TABLES
Table 4-35. Monthly Consumption Limits for Chronic Systemic Health Endpoints for
Children - Hexachlorobenzene
Chemical
Concentration
in Fish Tissue9
(mg/kg or ppm)
<0.04
0.04
0.05
0.06
0.07
0.08
0.09
0.1
0.2
0.3
0.4
0.5
0.6
0.7
0.8
0.9
1
2
3
4
5
6
7
8
>8
Recommended Risk - Based Consumption Limit (meals per month)"
3-oz Meal Size
(0.085 kq)
>30
>30
>30
>30
>30
>30
>30
>30
20
13
10
8
6
5
5
4
4
2
1
1
6/yr
6/yr
6/yr
6/yr
NONE
4-oz Meal Size
(0.11 4 kq)
>30
>30
>30
>30
>30
>30
>30
30
15
10
7
6
5
4 .
3
3
3
1
1
6/yr
6/yr
6/yr
NONE
NONE
NONE
8-oz Meal Size
(0.227 kq)
>30
>30
>30
25
22
19
17
15
7
5
3
3
2
2
1
1
1
6/yr
6/yr
NONE
NONE
' NONE
NONE
NONE
NONE
12-oz Meal Size
(0.341 kq)
>30
25
20
17
14
12
11
10
5
3
2
2
1
1
1
1
1
6/yr
, NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE = No consumption recommended.
6/yr = Consumption of no more than 6 meals per year is recommended.
> 30 = Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 meal per day).
a Instructions for modifying the variables in this table are found in Section 3.3.
Consumption limits are calculated based on a body weight of 14.5 kg and using RfD=8x10"4 mg/kg/d.
References for RfDs are found in Section 5.
The detection limit is 1 x 10~4 mg/kg.
All values were rounded down to the nearest whole meal size.
b Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD).
When this dose is consumed in less than 1 month (e.g., in a few large meals), the daily dose will
exceed the RfD (see Section 2.3).
4-39
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
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4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-37. Monthly Consumption Limits for Chronic Systemic Health Endpoints for the
General Population - Lindane
Chemical
Concentration
in Fish Tissue3
(mg/kgorppm)
<0.05
0.05
0.06
0.07
0.08
0.09
0.1
0.2
0.3
0.4
0.5
0.6
0.7
0.8
0.9
1
2
3
4
5
6
7
8
9
10
11
>11
Recommended Risk - Based Consumption Limit (meals per month)"
4-oz Meal Size
(0.11 4 kq)
>30
>30
>30
>30
>30
>30
>30
28
18
14
11
9
8
7
6
5
2
1
1
1
6/yr
6/yr
6/yr
6/yr
6/yr
6/yr
NONE
8-oz Meal Size
(0.227 kq)
>30
>30
>30
>30
>30
>30
28
14
9
7 .
5
4
4
3
3
2
1
6/yr
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE
12-ozMeal Size
(0.341 kq)
>30
>30
>30
26
23
20
18
9
6
4
3
3
2
2
2
1
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
16-ozMeal Size
(0.454 kq)
>30
28
23
20
17
15
14
7
4
3
2
2
2
1
1
1
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE = No consumption recommended.
6/yr = Consumption of no more than 6 meals per year is recommended.
> 30 = Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 meal per day).
a Instructions for modifying the variables in this table are found in Section 3.3.
Consumption limits are calculated based on an adult body weight of 70 kg and using RfD=3x10~4mg/kg/d.
References for RfDs are found in Section 5.
The detection limit is 1 x 10"4 mg/kg.
All values were rounded down to the nearest whole meal size.
Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD).
When this dose is consumed in less than 1 month (e.g., in a few large meals), the daily dose will
exceed the RfD (see Section 2.3).
4-41
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-38. Monthly Consumption Limits for Chronic Systemic Health Endpoints for
Children - Lindane
Chemical
Concentration
in Fish Tissue8
(mg/kg or ppm)
<0.02
0.02
0.03
0.04
0.05
0.06
0.07
0.08
0.09
0.1
0.2
0.3
0.4
0.5
0.6
0.7
0.8
0.9
1
2
3
>3
Recommended Risk - Based Consumption Limit (meals per month)b
3-oz Meal Size
(0.085 kg)
>30
>30
>30
>30
>30
25
22
19
17
15
7
5
3
3
2
2
1
1
1
6/yr
6/yr
NONE
4-oz Meal Size
(0.114kg)
>30
>30
>30
29
23
19
16
14
12
11
5
3
2
2
1
1
1
1
1
6/yr
NONE
NONE
8-oz Meal Size
(0.227 kq)
>30
29
19
14
11
9
8
7
6
5
2
1
1
1
6/yr
6/yr
6/yr
6/yr
6/yr
NONE
NONE
NONE
12-oz Meal Size
(0.341 kq)
>30
19
12
9
7
6
5
4
4
3
1
1
6/yr
6/yr
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE = No consumption recommended.
6/yr = Consumption of no more than 6 meals per year is recommended.
> 30 = Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 meal per day).
8 Instructions for modifying the variables in this table are found in Section 3.3.
Consumption limits are calculated based on a body weight of 14.5 kg and using RfD=3x10"4 mg/kg/d.
References for RfDs are found in Section 5.
The detection limit is 1 xlO"4mg/kg. ,
All values were rounded down to the nearest whole meal size.
b Monthly limits are based on the total dose allowable over a 1 -month period (based on the RfD).
When this dose is consumed in less than 1 month (e.g., in a few large meals), the daily dose will
exceed the RfD (see Section 2.3).
4-42
-------
4. RISK-BASED CONSUIWPTfON LflWfT TABLES
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4-43
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-40. Monthly Consumption Limits for Chronic Systemic Health Endpoints for the
General Population - Mirex
Chemical
Concentration
In Fish Tissue*
(mg/kg or ppm)
<0.04
0.04
0.05
0.06
0.07
0.08
0.09
0.1
0.2
0.3
0.4
0.5
0.6
0.7
0.8
0.9
1
2
3
4
5
6
7
>y
Recommended Risk - Based Consumption Limit (meals per month)"
4-oz Meal Size
(0.11 4 kq)
>30
>30
>30
>30
>30
>30
>30
>30
18
12
g
7
6
5
4
4
3
1
1
6/vr
6/yr
6/vr
6/yr
NONE
8-oz Meal Size
(0.227 kq)
>30
>30
>30
>30
26
23
20
18
9
6
4
3
3
2
2
2
1
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
12-oz Meal Size
(0.341 kq)
>30
>30
24
20
17
15
13
12
6
4
3
2
2
1
1
1
1
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
16-ozMeal Size
(0.454 kq)
>30
23
18
15
13
11
10
9
4
3
2
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE = No consumption recommended.
6/yr = Consumption of no more than 6 meals per year is recommended.
> 30 = Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 meal per day).
* Instructions for modifying the variables in this table are found in Section 3.3.
Consumption limits are calculated based on an adult body weight of 70 kg and using Rf0=2x10"" mg/kg/d.
References for RfDs are found in Section 5.
The detection limit is 1 x 10"4 mg/kg.
All values were rounded down to the nearest whole meal size.
b Monthly limits are based on the total dose allowable over a 1 -month period (based on the RfD).
When this dose is consumed in less than 1 month (e.g., in a few large meals), the daily dose will
exceed the RfD (see Section 2.3).
4-44
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-41. Monthly Consumption Limits for Chronic Systemic Health Endpoints for
Children - Mirex
Chemical
Concentration
in Fish Tissue3
(mg/kg or ppm)
<0.01
0.01
0.02
0.03
0.04
0.05
0.06
0.07
0.08
0.09
0.1
0.2
0.3
0.4
0.5
0.6
0.7
0.8
0.9
1
2
>2
Recommended Risk - Based Consumption Limit (meals per month)13
3-oz Meal Size
(0.085 kq)
>30
>30
>30
>30
25
20
17
14
12
11
10
5
3
2
2
1
1
1
1
1
6/yr
NONE
4-oz Meal Size
(0.114 kq)
>30
>30
>30
25
19
15
12
11
9
8
7
3
2
1
1
1
1
6/yr
6/yr
6/yr
NONE
NONE
8-oz Meal Size
(0.227 kq)
>30
>30
19
12
9
7
6
5
4
4
3
1
1
6/yr
6/yr
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
12-oz Meal Size
(0.341 kq)
>30
25
12
8
6
5
4
3
3
2
2
1
6/yr
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE = No consumption recommended.
6/yr = Consumption of no more than 6 meals per year is recommended.
> 30 = Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 meal per day).
a Instructions for modifying the variables in this table are found in Section 3.3.
Consumption limits are calculated based on a body weight of 14.5 kg and using RfD=2x10'4 mg/kg/d.
References for RfDs are found in Section 5.
The detection limit is 1 x 10"4 mg/kg.
All values were rounded down to the nearest whole meal size.
b Monthly limits are based on the total dose allowable over a 1 -month period (based on the RfD).
When this dose is consumed in less than 1 month (e.g., in a few large meals), the daily dose will
exceed the RfD (see Section 2.3).
4-45
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
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4-46
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-43. Monthly Consumption Limits for Chronic Systemic Health Endpoints for the
General Population - Toxaphene (OW/OPP RfD)
Concentration
in Fish Tissue8
(mg/kg or ppm)
0.06
0.07 ;:<"
0.08
0.09
0.1
0.2
0.3
0.4
0.5
0.6
0.7
0.8
0.9
1
2
3
4
5
6
7
8
9
10
11
12
13
>13
Recommended Risk - Based Consumption Limit (meals per month)13
4-oz Meal Size
(0.11 4 ka)
>30 . . -
>30- •
>30
>30
>30
>30
22
16
13
11
9
8
7
6
3
2
1
1
1
6/yr
6/yr
6/yr
6/yr
6/yr
6/yr
6/yr
NONE
8-oz Meal Size
(0.227 ka)
>30
>30
>30
>30
>30
>30
16
11
8
6
5
4
4
3
3
1
1
6/yr
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
12-oz Meal Size
>30
>30 , ,
>30 .
28
24
22
11
7
5
4
3
3
2
2
2
1
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
16-ozMeal Size
>30
28
24
21
18
16
8
5
4
3
2
2
2
1 .
1
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE = No consumption recommended.
6/yr = Consumption of no more than 6 meals per year is recommended.
> 30 = Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 -meal per day).
a Instructions for modifying the variables in this table are found in Section 3.3.
Consumption limits are calculated based on an adult body weight of 70 kg and using RfD=3.6x10"4 mg/kg/d
References for RfDs are found in Section 5.
The detection limit is 1 x 10"3 mg/kg.
All values were rounded down to the nearest whole meal size.
Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD).
When this dose is consumed in less than 1 month (e.g., in a few large meals), the daily dose will
exceed the RfD (see Section 2.3).
4-47
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-44. Monthly Consumption Limits for Chronic Systemic Health Endpoints for
Children - Toxaphene (OW/OPP RfD)
r==^=
Chemical
Concentration
In Fish Tissue8
(mg/kg or ppm)
— ^— =^
<0.02
0.02
0.03
0.04
0.05
0.06
0.07
0.08
0.09
0.1
0.2
0.3
0.4
0.5
0.6
0.7
0.8
0.9
1
2
3
Recommended Risk - Based Consumption Limit (meals per month)
3-oz Meal Size
(0.085 kg)
1
>30
>30
>30
>30
>30
>30
26
23
20
18
9
6
4
3
3
2
2
2
1
6/yr
6/yr
4-oz Meal Size
(0.11 4 kq)
>30
>30
>30
>30
27
23
19
17
15
13
6
4
3
2
2
1
1
1
1
6/yr
NONE
NONE
8-oz Meal Size
(0.227 kq)
>30
>30
23
17
13
11
9
8
7
6
3
2
1
1
1
6/yr
6/yr
6/yr
6/yr
NONE
NONE
NONE
12-oz Meal Size
(0.341 kg)
>30
23
15
11
9
7
6
5
5
4
2
1
1
6/yr
6/yr
6/yr
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE = No consumption recommended.
6/yr = Consumption of no more than 6 meals per year is recommended.
> 30 = Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 meal per day).
* Instructions for modifying the variables in this table are found in Section 3.3. ^
Consumption limits are calculated based on a body weight of 14.5 kg and using Rf 0=3.6x10" mg/kg/d.
References for RfDs are found in Section 5.
The detection limit is 1 x 10"3 mg/kg.
All values were rounded down to the nearest whole meal size.
b Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD).
When this dose is consumed in less than 1 month (e.g., in a few large meals), the daily dose will
exceed the RfD (see Section 2.3).
4-48
-------
4. RISK-BASED CONSUMPTION LflWfr TABLES
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-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-49. Monthly Consumption Limits for Chronic Systemic Health Endpoints for
Children - Diazinon
Concentration
in Fish Tissue8
(mg/kg or ppm)
0.004
0,005 " ,
0.006
0.007 "
- 0.008
0,009
0.01
0.02
0.03
0.04
0.05
0.06
0.07
0.08
,0.09
0.1
0.2
0.3
0.4
0.5
0.6
0.7
0.8
0.9
>0.9
Recommended Risk - Based Consumption Limit (meals per month)"
3-oz Meal Size
(0.085 kcri
>30
>30 ' .
>,30
, >30
>30
>30
" „ >30* -,
23
15
11
9
7
6
5
5
4
2
1
1
6/yr
6/yr
6/yr
6/yr
6/yr
NONE
4-oz Meal Size
(0.11 4 ka)
„ >30
,>30' -"
" - - »->30
>30
• . >30 -, -
"" .. >30
>30.
17
11
8
6
5
4
4
3
3
1
1
6/yr
6/yr
6/yr
NONE
NONE
NONE
NONE
8-oz Meal Size
>30
>30
>30
29
: - - 24 _
21
" 19
17 -
8
5
4
3
2
2
2
1
1
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
12-oz Meal Size
" >30
29
23' -• """
19 -
16
14 "-
12
'• 11
g
3
2
2
1
1
1
1
1
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE = No consumption recommended.
6/yr = Consumption of no more than 6 meals per year is recommended
> 30 = Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 meal per day).
a Instructions for modifying the variables in this table are found in Section 3.3.
Consumption limits are calculated based on a body weight of 14.5 kg and using RfD=9x10'5 mg/kq/d
References for RfDs are found in Section 5.
Shaded values are below the detection limit, 1 x 10~2 mg/kg.
All values were rounded down to the nearest whole meal size.
b Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD)
When this dose is consumed in less than 1 month (e.g., in a few large meals), the daily dose will
exceed the RfD (see Section 2.3).
4-53
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-50. Monthly Consumption Limits for Chronic Systemic Health Endpoints for the
General Population - Disulfoton
Chemical
Concentration
in Fish Tissue8
(mg/kg or ppm)
<0.007
0.007
0.008
0.009
0.01
0.02
0.03
0.04
0.05
0.06
0.07
0.08
0.09
0.1
0.2
0.3
0.4
0.5
0.6
0.7
0.8
0.9
1
Recommended Risk - Based Consumption Limit (meals per month)"
4-oz Meal Size
>30
>30
>30
>30
>30
>30
24
18
14
12
10
9
8
7
3
2
1
1
1
1
6/vr
6/vr
6/yr
8-oz Meal Size
(0.227 kq)
>30
>30
>30
>30
18
12
9
7
6
5
4
4
3
1
1
6/yr
6/yr
6/yr
6/yr
NONE
NONE
NONE
NONE
12-ozMeal Size
(0.341 kq)
>30
>30
27
24
12
8
6
4
4
3
3
2
2
1
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE
16-oz Meal Size
(0.454 kq)
26 '
23
20
, ,- 118
9
6
4
3
3
2
2
2
1
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE = No consumption recommended.
6/yr = Consumption of no more than 6 meals per year is recommended.
> 30 = Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 meal per day).
* Instructions for modifying the variables in this table are found in Section 3.3. ^
Consumption limits are calculated based on an adult body weight of 70 kg and using RfD=4x10" mg/kg/d.
References for RfDs are found in Section 5.
Shaded values are below the detection limit, 1 x 10~2 mg/kg.
All values were rounded down to the nearest whole meal size.
bMonthly limits are based on the total dose allowable over a 1-month period (based on the RfD).
When this dose is consumed in less than 1 month (e.g., in a few large meals), the daily dose will
exceed the RfD (see Section 2.3).
4-54
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-51. Monthly Consumption Limits for Chronic Systemic Health Endpoints for
Children - Disulfoton
Concentration
in Fish Tissue9
(mg/kg or ppm)
0,002 '
0.003 ,
0.004
*, >f 0,005- "
'" ^ - 0,006 . '
, 0.00?
, 0.008 _ "•
0.009 -
o.oV
0.02
0.03
0.04
0.05
0.06
0.07
0.08
0.09
0.1
0.2
0.3
0.4
>0.4
Recommended Risk - Based Consumption Limit (meals per month)*1
3-oz Meal Size
(0.085 kcrt
>30
>30*
>30
>30 „
* ^>30*
29 ,- -
25
23 • ;>
'20
10
6
5
4
3
2 '
2
2
2
1
6/yr
6/yr
NONE
4-oz Meal Size
(0.11 4 ka)
>30
>30
'* .- "„„ >30 * .
„ >.30 '
30
" ' „ 25 - k ,
22
19
.' ,.17-
15- - ""
7
5
3
3
2
2
1
1
1
6/yr
6/yr
NONE
NONE
8-oz Meal Size
, >30" -
>30
25
' 19
15
12
. - 11
9 '
3
- ' '' 7 ,
3
2
1
1
1
1
6/yr
6/yr
6/yr
NONE
NONE
NONE
12-oz Meal Size
>30
25
17 '
12
. 10
8 -
7
6
* 5
5
2
1
1
1
6/yr
6/yr
6/yr
6/yr
6/yr
NONE
NONE
NONE
NONE = No consumption recommended.
6/yr = Consumption of no more than 6 meals per year is recommended.
> 30 = Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 meal per day).
a Instructions for modifying the variables in this table are found in Section 3.3.
Consumption limits are calculated based on a body weight of 14.5 kg and using RfD=4x10'5 mg/kq/d
References for RfDs are found in Section 5.
Shaded values are below the detection limit, 1 x 10"2 mg/kg.
All values were rounded down to the nearest whole meal size.
Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD).
When this dose is consumed in less than 1 month (e.g., in a few large meals), the daily dose will
exceed the RfD (see Section 2.3).
4-55
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-52. Monthly Consumption Limits for Chronic Systemic Health Endpoints for the
General Population - Ethion
Chemical
Concentration
in Fish Tissue"
(mg/kg or ppm)
<0.08
0.08
0.09
0.1
0.2
0.3
0.4
0.5
0.6
0.7
0.8
0.9
1
2
3
4
5
6
7
8
9
10
15
Recommended Risk - Based Consumption Limit (meals per month)"
4-oz Meal Size
>30
>30
>30
>30
>30
>30
23
18
15
13
11
10
9
4
3
2
1
1
1
1
1
6/yr
6/yr
NONE
8-oz Meal Size
(0.227 kq)
>30
>30
>30
>30
23
15
11
9
7
6
5
5
4
2
1
1
6/yr
6/yr
6/yr
6/yr
6/yr
NONE
NONE
NONE
12-oz Meal Size
(0.341 kq)
>30
>30
>30
15
10
7
6
5
4
3
3
3
1
1
6/yr
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
16-ozMeal Size
(0.454 kq)
29
26
23
11
7
.5
4
3
3
2
2
2
1
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE = No consumption recommended.
6/yr = Consumption of no more than 6 meals per year is recommended.
> 30 = Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 meal per day).
8 Instructions for modifying the variables in this table are found in Section 3.3. •
Consumption limits are calculated based on an adult body weight of 70 kg and using RfD=5x10" mg/kg/d.
References for RfDs are found in Section 5.
The detection limit is 2 x 10"2 mg/kg.
All values were rounded down to the nearest whole meal size.
b Monthly limits are based on the total dose allowable over a 1 -month period (based on the RfD).
When this dose is consumed in less than 1 month (e.g., in a few large meals), the daily dose will
exceed the RfD (see Section 2.3).
4-56
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-53. Monthly Consumption Limits for Chronic Systemic Health Endpoints for
Children - Ethion
Chemical
Concentration
in Fish Tissue8
(mg/kg or ppm)
<0.03
0.03
0.04
0.05
0.06
0.07
0.08
0.09
0.1
0.2
0.3
0.4
0.5
0.6
0.7
0.8
0.9
1
2
3
4
5
>5
Recommended Risk - Based Consumption Limit (meals per month)"
3-oz Meal Size
(0.085 ka)
>30
>30
>30
>30
>30
>30
>30
28
25
12
8
6
5 • -
4
3
3
2
2
1
6/yr
6/yr
6/yr
NONE
4-oz Meal Size
(0.114ka)
>30
>30
>30
>30
>30
27
24
21
19
9
6
4
3
3
2
2
2
1
6/yr
6/yr
NONE
NONE
NONE
====^=
8-oz Meal Size
(0.227 kcrt
>30
>30
24
19
16
13
12
10
9
4
3
2
1
1
1
1
1
6/yr
NONE
NONE
NONE
NONE
NONE
==^=^=^=^=
12-oz Meal Size
(0.341 kcrt
>30
21
16
12
10
9
8
-7
6
3
2
1
1
1
6/yr
6/yr
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
^ -- 1 1,—.
NONE = No consumption recommended.
6/yr = Consumption of no more than 6 meals per year is recommended.
> 30 = Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 meal per day).
a Instructions for modifying the variables in this table are found in Section 3.3.
Consumption limits are calculated based on a body weight of 14.5 kg and using RfD=5x10'4mq/ka/d
References for RfDs are found in Section 5.
The detection limit is 2 x 10"2 mg/kg.
All values were rounded down to the nearest whole meal size.
"Monthly limits are based on the total dose allowable over a,1-month period (based on the RfD)
When this dose is consumed in less than 1 month (e.g., in a few large meals), the daily dose will
exceed the RfD (see Section 2.3).
4-57
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-54. Monthly Consumption Limits for Chronic Systemic Health Endpoints for the
General Population - Terbufos
Chemical
Concentration
in Fish Tissue"
(mg/kg or ppm)
<0.02
0.02
0.03
0.04
0.05
0.06
0.07
0.08
0.09
0.1
0.2
0.3
0.4
0.5
0.6
0.7
0.8
0.9
1
2
3
4
Recommended Risk - Based Consumption Limit (meals per month)"
4-oz Meal Size
>30
>30
>30
>30
>30
>30
>30
30
26
24
12
8
6
4
4
3
3
2
2
1
6/yr
6/vr
8-oz Meal Size
(0.227 kq)
>30
>30
>30
30
24
20
17
15
13
12
6
4
3
2
2
1
1
1
1
6/yr
NONE
NONE
NONE
12-oz Meal Size
(0.341 kq)
>30
27
20
16
13
11
10
9
8
4
2
2
1
1
1
1
6/yr
6/yr
NONE
NONE
NONE
NONE
16-oz Meal Size
(0.454 kq)
30
20 ,
15
12
10
8
7
6
6
3
2
1
1
1
6/yr
6/yr
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE = No consumption recommended.
6/yr = Consumption of no more than 6 meals per year is recommended.
> 30 = Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 meal per day).
" Instructions for modifying the variables in this table are found in Section 3.3. ^
Consumption limits are calculated based on an adult body weight of 70 kg and using Rf D=1.3x10 mg/kg/d.
References for RfDs are found in Section 5.
Shaded values are below the detection limit, 5 x 10"2 mg/kg.
All values were rounded down to the nearest whole meal size.
"Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD).
When this dose is consumed in less than 1 month (e.g., in a few large meals), the daily dose will
exceed the RfD (see Section 2.3).
4-58
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-55. Monthly Consumption Limits for Chronic Systemic Health Endpoints for
Children - Terbufos
Chemical
Concentration
in Fish Tissue3
(mg/kg or ppm)
<0.006
0,006
0.007
0.008
0.009
0,01
0.02
0.03
0.04
0.05
0.06
0.07
0.08
0.09
0.1
0.2
0.3
0.4
0.5
0.6
0.7
0.8
0.9
1
>1
Recommended Risk - Based Consumption Limit (meals per month)"
3-oz Meal Size
(0.085 ka)
'— a^ssi=ria=:^=^s=
>30
>30
>30
>30
;>30
>30
>30
22
'16
13.
I 11
9
8
7
6
3
2
1
1
1
6/yr
6/yr
6/yr
6/yr
NONE
4-oz Meal Size
(0.11 4 ka)
>30
>30
- >30 '.
- ->30
>30
>30
' 25 • -
. ;16
12
- 10
. 8
7
6
5
5
2
1
1
1
6/yr
6/yr
6/yr
6/yr
6/yr
NONE
8-oz Meal Size
(0.227 kcrt
r>30
• >30
>30
>30.
28
25
12
8
6
' -5
4
3
3
. 2
2
1
6/yr
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
12-oz Meal Size
(0.341 ken
^^ S^^SH
>30 I
28
24
- 21
18
16
8
4
3
2
2
2
1
1
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE = No consumption recommended.
6/yr = Consumption of no more than 6 meals per year is recommended.
> 30 = Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 meal per day).
Instructions for modifying the variables in this table are found in Section 3.3.
Consumption limits are calculated based on a body weight of 14.5 kg and using RfD=1 3x10-4mq/ka/d
References for RfDs are found in Section 5. '
Shaded values are below the detection limit, 5 x 10"2 mg/kg.
All values were rounded down to the nearest whole meal size.
b Monthly limits are based on the total dose allowable over a 1 -month period (based on the RfD)
When this dose is consumed in less than 1 month (e.g., in a few large meals), the daily dose will
exceed the RfD (see Section 2.3).
4-59
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-56. Monthly Consumption Limits for Chronic Systemic Health Endpoints for the
General Population - Oxyfluorfen
==SBS=S
Chemical
Concentration
in Fish Tissue*
(mg/kg or ppm)
-
30 [
>30
>30
>30
>30
>30
>30
28
18
14
11
9
8
7
6
5
2
1
1
1
6/yr
6/yr
6/yr
6/yr
6/yr
6/yr
8-oz Meal Size
(0.227 kg)
™" ""
>30
>30
>30
>30
>30
>30
28
14
9
7
5
4
4
3
3
2
1
6/yr
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE
12-oz Meal Size
(0.341 kg)
>30
>30
>30
26
23
20
18
9
6
4
3
3
2
2
2
1 '
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
16-oz Meal Size
(0.454 kg)
>30
28
23
20
17
15
14
7
4
3
2
2
1
1
1
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE = No consumption recommended.
6/yr = Consumption of no more than 6 meals per year is recommended.
> 30 = Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 meal per day).
* Instructions for modifying the variables in this table are found in Section 3.3. g
Consumption limits are calculated based on an adult body weight of 70 kg and using RfD=3x10' mg/kg/d.
References for RfDs are found in Section 5.
The detection limit is 1 x 10"2 mg/kg.
All values were rounded down to the nearest whole meal size.
b Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD).
When this dose is consumed in less than 1 month (e.g., in a few large meals), the daily dose will
exceed the RfD (see Section 2.3).
4-60
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-57. Monthly Consumption Limits for Chronic Systemic Health Endpoints for
Children - Oxyfluorfen
Chemical
Concentration
in Fish Tissue8
(mg/kg or ppm)
<0.2
0.2
0.3
0.4
0.5
0.6
0.7
0.8
0.9
1
2
3
4
5
6
7
8
9
10
20
30
>30
Recommended Risk - Based Consumption Limit (meals per month)"
3-oz Meal Size
(0.085 kg)
>30
>30
>30
>30
>30
25
22
19
17
15
7
5
3
3
2
2
1
1
1
6/yr
6/yr
NONE
4-oz Meal Size
(0.114kg)
>30
>30
>30
29
23
19
16
14
12
11
5
3
2
2
6/yr
NONE
NONE
8-oz Meal Size
(0.227 ka)
>30
29
. 19
14
11
9
8
7
6
5
2
1
1
1
6/yr
6/yr
6/yr
6/yr
6/yr
NONE
NONE
NONE
12-oz Meal Size
(0.341 ka)
^=^^s^==±^^=:^=
>30
19
12
9
7
6
c
4
4
3
1
1
6/yr
6/yr
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
^
NONE = No consumption recommended.
6/yr = Consumption of no more than 6 meals per year is recommended.
> 30 = Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 meal per day).
a Instructions for modifying the variables in this table are found in Section 3.3.
Consumption limits are calculated based on a body weight of 14.5 kg and using RfD=3x10'3 mq/ka/d
References for RfDs are found in Section 5.
The detection limit is 1 x 10"2 mg/kg.
All values were rounded down to the nearest whole meal size.
b Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD)
When this dose is consumed in less than 1 month (e.g., in a few large meals), the daily dose will
exceed the RfD (see Section 2.3).
4-61
-------
00 O,
•- !2
Q) O
-O I LO 8
s> .12
I1 Ł Ł -d o 1
D) « i± "g, CO t
Ł <& *? Z ~
x •;= ,- a> .P In
T- Ł o ^^ Ł
*j .— TJ Ł2 is S
'c (o «u _• .. 2
,_ ffi » o 2
c S 5 o o o
o .2 T-, ^ o 5
H 5^ o cd ^
g »
J!
C
^ Ł o fe -S
^ i « Q- S c
o «J !r c 3
!l
o —
••- c _
« .9 ns eo 5
> o "5. O) 8 g
C 3
0)
O
°§
B i 2? Ł -S
2 » = fe §
^ o 1 « =
-Ł O ro cc <
o
-------
4. RISK-BASED CONSUMPTfON UMIT TABLES
CO
CO O)
11
t5 „_
CD O
CO *,
s-g-
••—
W
ill
.22
Ł
•&-a
E § s
o5 -2 »
Q. m Ł
Ǥ
Ł 3 Ł
O CO -,
•*- H- O
tiSS
Mil
.« ^ CO Q.
.*; o w ^r
8 i
CD O
O "O
Ł•15
CB 13
O C
I a I S »I
S O Q. 05 g CO
sum
r^ — o O ft-
- CO -
Q. Q. Ł Ł
_
„ "*
^,
M UJ „ II
4-63
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-60. Monthly Consumption Limits for Chronic Systemic Health Endpoints for the
General Population - PCBs (Aroclor 1254)
Chemical
Concentration
in Pi«?h Tissue8
(mg/kg or ppm)
<0.004
0.004
0.005
0.006
0.007
0.008
0.009
0.01
0.02
0.03
0.04
0.05
0.06
0.07
O.Q8
009
01
0 2
0.3
0.4
0.5
0.6
0.7
>0.7
^T^^^"" "" '™ — ""
Recommended Risk - Based Consumption Limit (meals per month)"
4-oz Meal Size
f 0.1 14 kg)
>30
>3Q
>30
>30
>30
>30
>30
>30
18
12
9
7
6
5
4
4
3
1
1
6/yr
6/yr
6/yr
6/yr
NONE
is===^=^==
8-oz Meal Size
(0.227 kg)
1 -^^gg^^===^^^
>30
>30
>30
>30
26
23
20
18
9
6
4
3
3
2
2
2~
1
6/vr
6/yr
NONE
MPiMC
iNtJlNli:
NONE
NONE
NONE
'•— ^^g^^=^^=^
12-oz Meal Size
(0.341 kg)
>30
>30
24
20
17
15
13
12
6
4
3
2
2
1
1
r
1
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
i=======
16-oz Meal Size
(0.454 kg)
>30
___
• 3
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE = No consumption recommended.
6/yr = Consumption of no more than 6 meals per year is recommended.
> 30 = Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 meal per day).
" Instructions for modifying the variables in this table are found in Section 3.3. g
Consumption limits are calculated based on an adult body weight of 70 kg and using Rf D=2x10 mg/kg/d.
References for RfDs are found in Section 5.
Shaded values are below the detection limit, 5 x 10"2 mg/kg.
All values were rounded down to the nearest whole meal size.
"Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD)
When this dose is consumed in less than 1 month (e.g., in a few large meals), the daily dose will
exceed the RfD (see Section 2.3).
4-64
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-61. Monthly Consumption Limits for Developmental Health Endpoints for Women
of Reproductive Age - PCBs (Aroclor 1254)
Concentration
in Fish Tissue8
(mg/kg or ppm)
?« 0.004,
0,005
-0.006"'- "
0.007 * , --
- 0.008
0.009
. ,0:01 -
0.02 „ _
0.03
, • 0.04, V -
0.05'
0.06
0.07
0.08
0.09
0.1
0.2
0.3
0.4
0.5
0.6
0.7
>0.7
Recommended Risk - Based Consumption Limit (meals per month)"
4-oz Meal Size
(0.114 ka>
- *'"*- >30 ^
. ' >30- 1**'<
5j2*3Q J^ ^ — V^L
*V >3Q« -
'<- >30
* ~>3Q,
>30^ * ~^^
18
vJ2 \ ,- ,
9
7 it „ '
6
5
4
4
3
1
1
6/yr
6/yr
6/yr
6/yr
NONE
8-oz Meal Size
(0.227 ka)
>30
" k^>30, -
* ,. * / >sb - - '
" - >30 „
^ "_ ,26
- -23 -^
•?v, ,.j2Q
18
.""-^., 9
- * 6 -^ -,
*""-.. ' . 4
3 '
3
2
2
2
1
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
12-oz Meal Size
>30
' >30
24
' -" ,20
17 i
15 ""
" -* 13
12
- - 6
~ 4
3 -
2
2
1
1
1
1
6/yr
NONE
NONE
NONE
NONE
NONE
16-oz Meal Size
>30
23
w 18 " -
: is .. -•.
^,, '-IS-1*
-.. ^11 • .-
' , 10
- - - 9-
, V 4 '.
3
2
;.„ 1-
1
1
1
1
6/yr
NONE
NONE
NONE
NONE
NONE
NONE
NONE = No consumption recommended.
6/yr = Consumption of no more than 6 meals per year is recommended.
> 30 = Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 meal per day).
a Instructions for modifying the variables in this table are found in Section 3.3.
Consumption limits are calculated based on an adult body weight of 70 kg and using RfD=2x10"5mg/kg/d.
References for RfDs are found in Section 5.
Shaded values are below the detection limit, 5 x 10"2 mg/kg.
All values were rounded down to the nearest whole meal size.
b Monthly limits are based on the total dose allowable over a 1 -month period (based on the RfD).
When this dose is consumed in less than 1 month (e.g., in a few large meals), the daily dose will
exceed the RfD (see Section 2.3).
4-65
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-62. Monthly Consumption Limits for Developmental health Endpoints for Children
PCBs(Aroclor1254)
Chemical
Concentration
in Fish Tissue*
(mg/kg or ppm)
<0.0009
0.0009
0.001
0.002
0.003
0.004
0.005
0.006
0.007
0.008
0.009
0.01
0.02
0.03
0.04
0.05
0.06
0.07
0.08
0.09
0.1
0.2
Recommended Risk - Based Consumption Limit (meals per month)"3
3-oz Meal Size
>30
>30
>30
>30
>30
25
20
17
14
12
11
10
5
3
2
2
1
1
1
1
1
6/vr
4-oz Meal Size
(0.11 4 kq)
>30
>30
>30
25
19
15
12
11
9
8
' • 7
3
2
• : .1
1
1
1
6/yr
6/yr
6/yr
NONE
NONE
8-oz Meal Size
(0.227 ka)
>30
>30
. 19
12
9
7
6
5
4
4
3
1
1
6/yr
' 6/yr
6/yr
6/yr
NONE
NONE
NONE
NONE
NONE
12-oz Meal Size
(0.341 ka)
'- " 28 J*
„, 25
12 < ".
'8,"
-: 6
5
4>, " •, -
" XQ
"3 „ *
2
2
1"
6/yr
-^6/yr
6/vr
NONE
NONE
NONE
NONE
NONE
NONE
NONE
NONE = No consumption recommended.
6/yr = Consumption of no more than 6 meals per year is recommended.
> 30 = Although consumption of more than 30 meals/month is allowed, EPA advises limiting
consumption to 30 meals in 1 month (1 meal per day).
* Instructions for modifying the variables in this table are found in Section 3.3. g
Consumption limits are calculated based on a body weight of 14.5 kg and using RfD=2x10" mg/kg/d.
References for RfDs are found in Section 5.
Shaded values are below the detection limit, 5 x 10 "2 mg/kg.
All values were rounded down to the nearest whole meal size.
b Monthly limits are based on the total dose allowable over a 1-month period (based on the RfD).
When this dose is consumed in less than 1 month (e.g., in a few large meals), the daily dose will
exceed the RfD (see Section 2.3).
4-66
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
Table 4-63. Monthly Consumption Limits for Carcinogenic Health Endpoints
for the General Population - PCBs (Total)
ecommended Risk - Based Consumption Limit (meals per month)"
cc
1
^
L
C
a
V
1
i
v
c
<7
"a
a
IV
t\
1 8-oz Meal Size (0.227 kg)
Ł
J5
5
i)
S
8
c
i
c
5
Chemical
Concentration
in Fish Tissue3
(mg/kg or ppm)
<° O "• ' " *'" • i> i. UJ 1
T- A "^ "^ *". "J m ^ ^ r" »««r|t!
T- AAAAAA'AA,'"
•"* "it '
T- A A, A A A' A A'™ A"™*;
M?f^s_^
p- AA''AA™"
-------
4. RISK-BASED CONSUMPTION LIMIT TABLES
3-
i found in Section 3.3.
Ddy weight of 70 kg
.2
H
E
0
T™
ow detection limit, 5 x
CD
-Q
CD
CO
CO
CD
"i
1 Shaded vali
S
ency factor of 1.56.x 1
•s
Q-
L.
potency factors are fo
d down to the nearest
L«
I
Ł
and using a
References
0)
I
1
CD
0)
All values w
f
. T3
1 8.
^ «
CO « -
co c
i =
t 0
T3 E
CO T-
1
CO
E
CD
•2
f
CD
1
II
_J
Lr
•d
^3
c
insumption recomme
o
o
o
II
111
Z
O
Z
C
CO
Ł
CD
0
O
C
•5
c:
o
"a.
E
CO
c
o
O
•^
CD
"S
•a
c
per year is recomme
_co
"7S
CD
CO
1
CD
Ł
h consumption of me
i1
O
Ł
n
0
CO
A
0)
O
CO
o
miting consumption t
CO
CD
CO
'•s
CO
^
Q.
LU
4-68
-------
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 (OPP)
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. Although EPA
has developed guidelines for study selection, it is clear that for many chemicals a
number of study results could be used to estimate exposure limits. The reader is
urged to review the information presented, particularly the studies of chemicals of
interest in their areas, so that they may choose the optimal health endpoints from
among those discussed in this document (e.g., carcinogenic toxicity, chronic
exposure toxicity) or develop their own risk values, based upon their review of the
information.
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-
5-1
-------
5.1 INTRODUCTION
cinogenic) effects. These include pharmacokinetics, acute and chronic toxicity,
developmental toxicity, mutagenicity, carcinogenicity, special susceptibilities, inter-
active 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 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, obtained primarily from ATSDR toxicological profiles, 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.
5-2
-------
5.1 INTRODUCTION
Table 5-1. Health and Toxicological Data Reviewed for Target Analytes
Category Specific Information
Background
Pharmacokinetics
Acute toxicity
Chronic toxicity
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
5-3
-------
5.1 INTRODUCTION
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. In
addition, the Minimum Risk Levels developed by ATSDR are included when
available. They provide estimates of the levels of exposure for a chemical (e.g.,
toxaphene) at which minimum risk is expected to occur (ATSDR, 1990b). In
addition, Appendix C contains a discussion of general class information for two
major categories of chemicals, the organochlorines and organophosphates, which
constitute 14 of the 25 target analytes.
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 includes 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 unusual aspects of the 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 Developmental Toxicity—
Developmental toxicity data were obtained for each target analyte (dioxin
information will be provided when the EPA dioxin reassessment is complete in
1998). Section 2.3.2.3 contains general information on developmental toxicity,
including definitions, methods for calculating exposure limits, and special issues
related to developmental toxicity. The data and methods information are provided
' to give readers the option of developing exposure limits based on developmental
effects, as they deem necessary.
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5.1 INTRODUCTION
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. If a female has been exposed to endrin, 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.
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,
all results could not be feasibly listed. 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 current 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, 1997).
For a complete description of EPA's current weight-of-evidence classification
scheme, see EPA's Guidelines for Carcinogenic Risk Assessment (U.S. EPA,
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5.1 INTRODUCTION
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, 1996d).
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 to exposure
to some chemicals. Qualitative data on special susceptibilities are provided for
many of the target analytes. In addition, information is provided on susceptibilities
of special concern for groups of chemicals (e.g., organophosphates) in Appendix
C. 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. The U.S. Public Health Service has provided specific nonquanti-
tative guidance regarding susceptible subgroups in the ATSDR Toxicity Profiles;
it is included in the individual toxicological profiles in Sections 5.2 through 5.8. In
addition, there are some general caveats regarding special susceptibilities that
should be considered. Exposure to many types of toxicants poses higher risks to
children due to their immaturity:
embryos, fetuses, and neonates up to age 2 to 3 months may be at
increased risk of adverse effects ... because their enzyme detoxifi-
cation systems are immature
and
Infants and children are especially susceptible to immunosuppres-
sion because their immune systems do not reach maturity until 10
to 12 years of age (ATSDR, 1990b).
ATSDR has also cautioned that: :
the elderly with declining organ function and the youngest of the
population with immature and developing organs will generally be
more vulnerable to toxic substances than healthy adults (ATSDR,
1993a).
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5.1 INTRODUCTION
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 Toxicological Profiles.
Often the data indicate that certain classes of chemicals may be of concern. For
example, most organochlorines induce the mixed function oxidase system. These
chemicals may lead to unanticipated and exaggerated or diminished 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. For example, if very limited study data
are available on developmental toxicity, but developmental toxicity is indicated in
the database, developmental studies are listed as a data gap.
5.1.1.10 Summary of EPA Levels of Concern—
The EPA risk values (RfDs and cancer slope factor) discussed in each section and
used in the development of fish consumption limits are summarized in Table 3-1.
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 1997 for cancer .slope factor, chronic
exposure RfDs, and additional study data. ATSDR Toxicological Profiles were also
consulted when available. The profiles have extensive toxicity, pharmacokinetic,
and epidemiological data reviews and provide estimated Minimum Risk Levels,
which are analogous to RfDs in that they are "estimates of levels posing minimal
risk to humans" (ATSDR, 1992a). They are based upon risk assessment methods
similar to those used by EPA. The ATSDR documents were particularly useful
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5.1 INTRODUCTION
because they provide detailed information and because many provide extensive
discussions of developmental effects as well as some MRLs for these effects.
Some ATSDR profiles cited are draft documents; however, the profiles underwent
extensive review within and outside of the U.S. Public Health Service before they
were released as the draft bound copies that are cited in this work.
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 considerable
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 few data, only a qualitative description may be possible.
5.1.2 Abbreviations Used and Scientific Notation
The abbreviations NOEL, NOAEL, LEL, LOEL, and LOAEL are used in this docu-
ment as they appear in the original sources. Although they have specific meanings
(see the Glossary), NOEL-NOAEL and LEL-LOEL-LOAEL are sometimes used
interchangeably. Since it was not possible to determine the intent of the authors
of the source documents, the terms were used as they appeared in those
documents.
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5.1 INTRODUCTION
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.
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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,
1993e).
5.2.1.2 Pharmacokinetics—
Pharmacokinetic studies show that water-soluble arsenic 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, 1993e).
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 metabolitis is in the urine, with human studies
showing that 45 to 85 percent is excreted in the urine within 1 to 3 days. Very little
is excreted in the feces (ATSDR, 1993e).
5.2.1.3 Acute Toxicity—
Arsenicals have been recognized as a human poison since ancient times, and
large doses, approximately 600 ug/kg/d or higher, taken orally have resulted in
death. Oral exposure to lower levels of arsenic has resulted in effects on the
gastrointestinal system (nausea, vomiting); central nervous system (headaches,
weakness, delirium); cardiovascular system (hypotension, shock); and the liver,
kidney, and blood (anemia, leukopenia). Because significant information is
available on the acute effects of arsenic poisoning in humans, few animal studies
have been carried out. The limited available data have shown arsenic to have low
to moderate acute toxicity to animals. This is based on data showing the LD50s for
arsenic to range between 50 and 5,000 mg/kg (ATSDR, 1993e).
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
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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 (ATSDR, 1993e).
Other effects noted from chronic oral exposure include peripheral neuropathy,
cardiovascular 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
- art uncertainty factor of 3. This was based on two studies that showed that the
prevalence of blackfoot disease increased with both age and dose for individuals
exposed to high levels of arsenic in drinking water. This same population also
displayed a greater incidence of hyperpigmentation and skin lesions. 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 uncertainly
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, 1997). ATSDR has
calculated a chronic oral MRL that is equal to the RfD listed in IRIS (ATSDR
1993e).
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 effects
in a large number of people. However, doses were not well-characterized, other
contaminants 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, 1997).
5.2.1.5 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. However, due to the small
number of cases, this association might be due to random variation. In another
study, a marginal association (not statistically significant) was found between
detectable levels of arsenic in drinking water and spontaneous abortions.
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, 1993e).
Minimal or no effects on fetal development have been observed in studies on
chronic oral exposure of pregnant rats or mice to low levels of arsenic in drinking
water. Malformations were produced in 15-day hamster fetuses via intravenous
injections of arsenic into pregnant dams on day 8 of gestation, while another study
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5.2 METALS
reported that very high single oral doses of arsenic were necessary to cause
prenatal fetal toxicity (IRIS, 1997).
5.2.1.6 Mutagenicity—•
In vivo studies of arsenic have shown mixed results. Some studies on
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, 1993e). In vitro studies
have also reported both positive and negative results. Arsenic was negative in the
bacterial colorimetric assay: SAS Chromotest (HSDB, 1997), and positive for
reverse mutations in bacteria, morphological transformations in Syrian hamster
embryo cells, and chromosomal aberrations in human leukocytes (ATSDR,
1993e).
5.2.1.7 Carcinogenicity—
There is clear evidence that chronic exposure of humans to inorganic arsenic
increases the risk of cancer. Ingestion of arsenic has been associated with an
increased risk of nonmelanoma skin cancer, and bladder, liver, and lung cancer.
In addition, studies have reported that inhalation of arsenic results in an increased
risk of lung cancer (IRIS, 1997).
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; the mechanism of action in
causing human cancer is not known, and rodents may not be a good model for
arsenic-induced carcinogenicity (IRIS, 1997).
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 skin, bladder, liver, and lung cancer through
drinking water exposure (IRIS, 1997).
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, 40,421 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, 1997).
5.2.1.8 Special Susceptibilities—
ATSDR reported that no studies were located regarding unusual susceptibility of
any human subpopulation to arsenic. However, it is possible that some members
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5.2 METALS
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, 1993e; 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.
Suggestive evidence exists 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, 1993e)
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, such as short-
term animal studies to define an acute or intermediate-duration MRL In addition,
epidemiological studies to provide additional support for the threshold dose for
arsenic in humans are lacking and would be valuable. Additional studies on
developmental and reproductive effects of arsenic would also be useful (ATSDR,
1993e).
5.2.1.11 Summary of EPA Levels of Concern—
Chronic Toxicity 3.0 x 10"4mg/kg/d
Carcinogenicity 1.5 per mg/kg/d.
5.2.1.12 Major Sources—
ATSDR (1993e), HSDB (1997), IRIS (1997), Rogers (1995).
5.2.2 Cadmium
5.2.2.1 Background—
Cadmium is a heavy metal that is released 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, 1993a).
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
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5.2 METALS
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 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, and this may
be due to the particular forms of cadmium in their food (ATSDR, 1993a).
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, 1993a).
Cadmium absorption appears to involve sequestering 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, 1993a). Once cadmium is absorbed, it is eliminated slowly; the
biological half-life has been estimated at 10 to 30 years (U.S. FDA, 1993).
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, 1993a).
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 humans, lethal doses caused
massive fluid loss, edema, and widespread organ destruction. The ingested doses
were 25 mg/kg and 1,500 mg/kg (ATSDR, 1993a).
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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, 1993a). 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 0.005 mg/kg/d in multiple human studies
of waterborne cadmium. 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 (IRIS, 1993).
The FDA has calculated a tolerable daily intake of 55 ug/person/day, 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 FDA (1993) for more details.
ATSDR has also recently calculated a risk value for oral exposure based on kidney
toxicity in humans. They developed a chronic MRL of 7 x 10~4 mg/kg/d based on
a NOAEL of 0.0021 mg/kg/d in a large human cohort. The critical endpoint was an
elevation of urinary beta-(2)-microglobulin. A toxicity threshold was estimated
using a kinetic model of cadmium metabolism that predicted that approximately 5
percent of nonsmokers will reach or exceed the dose required to cause an effect
at the NOAEL. To calculate the MRL, ATSDR used an addition uncertainty factor
of 3 to account for sensitive members of the population. However, the critical study
used a large population that included the elderly, who are considered a sensitive
subpopulation (ATSDR, 1993a). The MRL developed by ATSDR is within 1 order
of magnitude of the RfD developed by IRIS.
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 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 has been
suggested in some human studies; however, the results are conflicting (ATSDR
1993a).
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Animal studies indicate that cadmium 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 0.57 mg/kg/d in a 10-week study in
mice (ATSDR, 1993a). No longer-term studies were located for this work. An
alternative exposure could be calculated for immunological effects based on the
above study. The standard uncertainty factors used in this calculation would
typically take into consideration inter- and intraspecies variability, use of a less
than lifetime study, and the use of a LOAEL rather than a NOAEL. Immunological
effects require further investigation to determine whether this is an effect that
occurs in humans. It appears to be a sensitive endpoint for chronic exposure
toxicity.
5.2.2.5 Developmental Toxicity—
Developmental toxicity has been associated with cadmium exposure both in short-
and long-term studies. In 10-day prenatal dosing studies in rats at 18.4 mg/kg,
malformations including split palate and dysplasia of the facial bones 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 weight in the range of 2 to 20 mg/kg/d (ATSDR, 1993a). Oral cadmium
exposure of young mice depresses their humoral immune responses; the study did
not find the same effect in adult mice (ATSDR, 1993a).
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, which was
impaired in offspring of female rats orally exposed to cadmium at a
dose of 0.04 mg/kg/day prior to and during gestation ... (ATSDR,
1993a).
Reduced locomotor activity and impaired balance were noted at a LOEL 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 studies yielding similar
results were conducted with maternal exposures of 4.3 to 17.2 ug/mL of water (see
numerous citations in Baranski et al., 1983).
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
lacking statistical significance. Inhalation studies in animals have found structural
and neurobehavioral abnormalities similar to those found in the oral dosing studies
(ATSDR, 1993a).
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Based on the mutagenieity data results (discussed below), heritable defects may
result from exposure to cadmium. However, mutagenieity 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, 1993a).
Based on the reviewed information, neurobehavioral effects appear to be a critical
endpoint for developmental effects as indicated by the LOEL of 0.04 mg/kg/d. The
standard uncertainty factors used in the calculation of an exposure limit would
typically take into consideration inter- and intraspecies variability and the use of a
LOEL rather than a NOAEL.
Estimating an exposure limit for cadmium based on developmental toxicity is
problematic because the average daily dose is approximately 0.03 mg/d (ATSDR,
1993a), which is equivalent to 4 x 10"4per mg/kg/d in a 70-kg individual. The
exposure for developmental effects, which would be calculated using the
: neurobehavioral LOEL noted above (approximately 4 x 10'5 mg/kg/d), is one-tenth
of the average background consumption rate. Due to the margin of safety
introduced by these factors, the estimated exposure limit should be viewed in the
context of the overall exposure of population groups from all sources, as well as
the benefits of fish consumption. Balancing risks and benefits is discussed in
Volume 3 in this series, Risk Management,
5.2.2.6 Mutagenieity—
Results of bacteria, yeast, and human lymphocyte assays have been mixed.
Positive results were observed 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 aneupioidy. Positive results have also been obtained in
Chinese hamster ovary and mouse lymphoma cell assays (IRIS, 1993).
5.2.2.7 Carcinogenicity—
No animal or human oral exposure studies suggest that cadmium is carcinogenic
via the oral exposure route. Animal studies conducted at relatively low exposure
levels (up to 4.4 mg/kg/d) have yielded negative results. Studies have been
conducted on population groups in high cadmium,exposure areas 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. ATSDR concluded that there was little evidence of an association
between cadmium exposure and increased cancer risk in humans but that the
statistical power of the studies to detect an effect was not high. They determined
that neither the human nor animal studies provided sufficient evidence to
determine the carcinogenic status of cadmium (ATSDR, 1993a).
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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"3per ug/
m3 (IRIS, 1993).
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. As stated
above, 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, 1993a).
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 wprpen 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. Animal studies have
found an association between lindane and increased developmental toxicity and
between calcium deficiency and increased fetotoxicity. Ethanol increased liver
toxicity and garlic decreased kidney toxicity. Lead increased neurotoxicity and
selenium decreased the clastogenic effect of cadmium on bone marrow. Exposure
to chemicals that induce metaliothionein (e.g., metals) reduced toxicity with
parenteral cadmium exposure (ATSDR, 1993a).
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
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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, 1993a).
5.2.2.11 Summary of EPA Levels of Concern—
Chronic Toxicity
Carcinogenicity
5.2.2.12 Major Sources—
1 x 10"3mg/kg/d
Probable inhalation carcinogen (B1). Insufficient data to
determine carcinogenic status via oral exposure route.
ATSDR (1993a), HSDB (1993), IRIS (1993), 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 of fish 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. 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 extensively peer reviewed including a
recent review by the Science Advisory Board (SAB). At this time, the Mercury
Study Report to Congress has not been released as final. The SAB review draft
is available from NTIS.
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, 1991,1993a; Stern, 1993).
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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, 1996a).
5.2.3.2 Pharmacokinetics—
Methylmercury is rapidly and nearly completely absorbed; EPA and ATSDR have
used an estimate of 95 percent absorption following oral exposure (ATSDR, 1994;
U.S. EPA, 1996e), and the World Health Organization (WHO) has similarly
estimated an absorption of 90 to 100 percent for methylmercury (WHO, 1990).
Methylmercury is lipophilic, allowing it to pass through lipid membranes of cells
and facilitating its distribution to all tissues, following absorption from the
gastrointestinal tract. Methylmercury also binds readily to proteins. Methylmercury
is found throughout fish tissue, and a substantial portion of the mercury in fish can
be found in trimmed filets. Because of this, methylmercury exposure is not
significantly reduced by trimming fat and skin from fish prior to cooking.
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
demethylated to form mercuric mercury. In experiments on animals, females
eliminated mercury more slowly than males, and young animals more slowly than
adults. Neonatal excretion is slowed by the immaturity of the transport system.
Methylmercury readily crosses the placental and blood/ brain barriers. Estimates
for the hajf-life of methylmercury range from 44 to 80 days (U.S. EPA, 1996).
Excretion of methylmercury is via the feces, urine, and breast milk. Methylmercury
is 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 kidney damage and
failure, gastrointestinal damage, cardiovascular collapse, shock, and death. The
estimated lethal dose is 10 to 60 mg/kg (ATSDR, 1994). An acute/ intermediate
oral MRL of 1.2 x 10'4was calculated by ATSDR using the Iraqi data on in utero
exposed children described in Section 5.2.3.4.
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 so
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whether exposure occurs to the developing embryo or fetus during pregnancy or
to adults and children.
Exposure of humans 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 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; it accumulated in the tissue of
fish and shellfish that were a routine part of the diet in these populations. Average
fish consumption was reported to be in excess of 300 g/d (reviewed by Harada et
al., 1995); this is a level of fish consumption that is almost 50 .times greater than
is typical (6.5 g/d) for the general U.S. population.
Symptoms of Minamata disease in children and adults included the following:
impairment of the 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,
impairment of hearing, impairment of walking, and mental disturbances. It
sometimes took several years before people were aware that they were
developing the signs and symptoms of methylmercury poisoning. Over the years,
it became recognized 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. As in Minamata, multiple chemical plant sources of the chemical were
considered. The signs and symptoms of disease in Niigata were those of
methylmercury poisoning and the disease in Minamata.
Methylmercury poisoning 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 outbreak of methylmercury poisoning from grain
consumption occurred in the early 1970s. Imported mercury-treated seed grains
arrived after the planting season; the grain was 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. Because many of the people exposed to
methylmercury in this way lived in small villages in very rural areas (and some
were nomads), the total number of people exposed to these mercury-contaminated
seed grains is not known. The number of people admitted to the hospital with
symptoms of poisoning has been estimated to be approximately 6,500, with 459
fatalities reported.
As in the Japanese poisoning incidents, the signs and symptoms of disease were
predominantly in the nervous system: difficulty with peripheral vision or blindness,
sensory disturbances, incoordination, impairment of walking, slurred speech, and,
in some cases, death. Children were affected as well as adults. Of great concern
was the observation that infants, born of mothers who had consumed the
methylmercury-contaminated grain (particularly during the second trimester of
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pregnancy), could show nervous system damage even though the mother was
only slightly affected herself.
More recent studies have dealt with populations that are expected to be exposed
to methylmercury as a consequence of routine consumption of fish and marine
mammals. These have included studies of populations around the Great Lakes,
in New Zealand (Kjellstrom et al., 1986a, b), in the Amazon basin (e.g., Lebel et
al., 1996; Marsh et al., 1995), the Seychelles Islands (Marsh et al., 1995), and the
Faroe Islands (Dahl et al., 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, at the time of this writing,
still being analyzed and published.
Methylmercury health endpoints other than neurotoxicity were evaluated by EPA
using established risk assessment guidelines. Data for endpoints other than
developmental neurotoxicity were limited (see Section 5.2.3.5).
In 1985 EPA published an RfD for methylmercury in IRIS of 3 x 10"4mg/kg/d. The
critical effect was multiple central nervous system effects (including ataxia and
paresthesia) in adults in the Iraqi population who had been exposed to
methylmercury through consumption of contaminated grain (Clarkson et al., 1975).
A LOAEL of 0.003 mg/kg/d (corresponding to a blood concentration of 200 ug/L)
was determined from inspection of the data. An uncertainty factor of 10 was
applied for the use of a LOAEL in the absence of a NOAEL Since that time, EPA
has received several critiques and submissions to IRIS that questioned whether
this RfD, based on effects in adults, was protective against developmental effects.
A reexamination of the RfD took place with consensus on a revised value, and the
RfD became available on IRIS in May 1995. The basis, derivation, and uncertainty
analysis of the current EPA RfD is described at length in Volume IV, Health Effects
of Mercury and Mercury Compounds, Mercury Study Report to Congress (U.S.
EPA, 1996a).
The current EPA RfD for methylmercury was 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
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
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• 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 RfD for methylmercury is 1 x 10~4 mg/kg/d or 0.1 ug/kg/d.
The range of uncertainty in the 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 no longer the basis for EPA's
methylmercury RfD.
There are, however, uncertainties associated with the current 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 RfD is supported by additional studies in children exposed in utero. These
include investigations among Cree Indians in Canada and New Zealanders who
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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 (data on children of about 5
years of age) are also consistent with EPA's benchmark dose.
A recent review by the Science Advisory Board (SAB) determined that, at this time,
the RfD based on the data on Iraqi children 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.
All RfDs are defined as having a degree of uncertainty of perhaps an order of
magnitude. The RfD may be considered to be the midpoint in an estimated range
of an order of magnitude (a factor of 10 or two factors of 3). Assuming that the RfD
of 1 x 10'4 mg/kg/d is the midpoint of an order of magnitude range of uncertainty,
then the upper end of the RfD range is 3 x 10'4 mg/kg/d and the lower bound value
is 7 x lO'^mg/kg/d. It is useful to estimate the number of fish meals per week that
would result in exposure at the RfD. For a 70-kg person, the RfD is 7 x 10" mg/d
or 4 9 x 10'2 mg/wk. If one assumes an 8-oz (0.227-kg) meal size and a fish tissue
contamination level of 0.2 ppm (0.2 mg/kg), then one fish meal per week would
result in exposure at the RfD. Given that there is a range threefold above and
below the point estimate of the RfD, this consumption limit spans as many as three
fish meals per week or as few as one fish meal every 3 weeks. Calculation of fish
tissue contamination ranges for one 8-oz fish meal per week using methylmercury
as an example is provided in Section 3.2.2.1.
5.2.3.5 Developmental Toxicity—
There are data linking elemental, mercuric, and methylmercury with developmental
effects; these were used by EPA to determine weight-of-evidence classifications
as specified in the Guidelines for Risk Assessment of Developmental Toxicants.
For methylmercury, there are data on developmental effects in rats, mice, guinea
pigs, hamsters, and monkeys. As described above (and documented at length in
Volume IV of the Mercury Study Report to Congress [U.S. EPA, 1996a]), there are
convincing data from a number of human studies that methylmercury is a
developmental toxicant resulting in subtle to severe neurologic effects depending
on dose and individual susceptibility. According to EPA guidelines, methylmercury
is classified as having sufficient human and animal data for developmental toxicity.
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. As a result of this, it is advisable to
reduce mercury exposure of girls and women with childbearing potential to reduce
overall body burden. If a woman has been exposed to mercury, even if exposure
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is reduced during pregnancy, the outcome of that pregnancy may be affected,
depending on the timing and extent of prior exposure.
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 DNA. In
humans, methylmercury is widely distributed in the body. There are data on
animals indicating that methylmercury administered intraperitoneally reaches germ
cells and may produce adverse effects in those cells. Sex-linked recessive
mutations (a sign of chromosomal damage to germ cells) were increased in
Drosophila melanogaster given methylmercury in the diet. Studies have reported
increased incidence of chromosomal aberrations (Skerfving et al., 1970) or sister
chromatid exchange (Wulf et al., 1986) in lymphocytes of humans ingesting
mercury-contaminated fish or meat. Chromosome aberrations have been reported
in cats treated in vivo and in cultured human lymphocytes in vitro. Evidence of
DNA damage has been shown in a number of in vitro systems.
Using criteria in the Guidelines for Mutagenicity Risk Assessment (U.S. EPA,
1986c), the EPA has classified methylmercury as being of high concern for
potential human germ cell mutagenicity. All that keeps methylmercury from the
highest level of concern is the lack of positive results in a heritable mutation assay.
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. Dietary exposures of mice to methylmercury resulted in significant
increases in the incidences of kidney tumors in males but not in females (U.S.
EPA, 1996e). EPA has classified methylmercury as a Group C, possible human
carcinogen, based on inadequate data in humans and limited evidence in animals.
EPA has not calculated quantitative carcinogenic risk values for methylmercury
(IRIS, 1997). It should be noted that 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. The data from genotoxicity testing
indicate that, although methylmercury is clastogenic (breaks chromosomes), it
does not cause point mutations. Evidence points to a mode of action for methyl-
mercury 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 of fish, methylmercury is not
likely to present a carcinogenic risk to the U.S. population.
5.2.3.8 Special Susceptibilities—
The developing fetus is thought to be at increased risk from methylmercury
exposure. There are not sufficient data on children exposed only after birth to
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determine if this is a group with increased susceptibility to mercury toxicity.
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 adults. Additional risk may also result from the apparently decreased ability of
young individuals to eliminate mercury (see Section 5.2.3.2). ATSDR has listed the
following groups as particularly susceptible: people with impaired organ function
(especially kidney, CMS, and liver) and individuals with a dietary insufficiency of
zinc, glutathione, antioxidants, or selenium (ATSDR, 1994).
5.2.3.9 Interactive Effects-
Potassium dichromate and ethanol may increase the toxicity of mercury, although
these effects have been noted only with metallic and inorganic mercury. Atrazine
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, 1994). There is insufficient information to
recommend quantitative changes in risk estimations based upon interactive
effects.
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 Levels of Concern-
Chronic Toxicity
Carcinogenicity
Developmental Toxicity
1 x1Q-4mg/kg/d
Insufficient data to determine carcinogenic status.
No developmental risk value calculated; the chronic
toxicity RfD above was determined based on devel-
opmental effects.
5.2.3.12 Major Sources—
ATSDR (1994), IRIS (1995, 1997), Shubat (1993a); Stern (1993), U.S. EPA
' (19933,1995,1996a). '".'.'.'.'''
5.2.4 Selenium
5.2.4.1 Background-
Selenium 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 nonlactating women and 20 additional ug/d during
lactation ATSDR has identified daily intake at nontoxic levels of approximately
0 05 to 0.15 mg/d (ATSDR, 1989; HSDB, 1993). This is approximately equivalent
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to 7 x 1(r4 to 2 x 10"3 mg/kg/d in a 70-kg individual. The RDA for adult males is 70
ug/d (NRC, 1989). 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. It has a|so been shown to have a protective effect against
chemically induced cancers in laboratory animals (Bobbins et al., 1989). Although
selenium is an essential nutrient, it is toxic at high exposure levels and is
mutagenic in some test systems (ATSDR, 1989).
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, 1989). 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 (ATSDR, 1989). Detailed information on metabolism of selenium
can be found in the Toxicological Profile for Selenium. This document also
contains an extensive discussion of the selenium concentrations in human tissues
and fluids correlated with specific health effects (ATSDR, 1989).
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 (IRIS, 1993). Subacute selenosis
(prolonged exposure at relatively high doses) causes impaired vision, ataxia,
disorientation, and respiratory distress (IRIS, 1993). Tachycardia has beeri
reported in humans exposed to high doses; myocardial disorders have also been
associated with selenium deficiencies. Acute exposure dog studies at high doses
have found multiple alterations in blood chemistry (ATSDR, 1989).
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 (IRIS,
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1993) Note that the NOAEL and LOAEL for selenium in the 1989 human study are
only slightly higher than the average daily intake and the RDA (see above).
High levels of selenium exposure have caused the following effects: lowered
hemoglobin levels, mottled teeth, skin lesions, CNS abnormalities, fatigue
anorexia, enlarged spleen, thickened and brittle nails, hair and nail loss, decreased
blood clotting ability, liver dysfunction, and muscle twitching (IRIS, 1993). Humans
exposed to high dietary levels have reported Gl disturbances (dose unspecified).
Cows with high naturally occurring dietary exposures were found to have ulcers
in the upper Gl tract (ATSDR, 1989).
Lifetime exposure of mice to sodium selenate or sodium selenite at 0.31 mg/kg/d
caused amyloidosis of the lung, liver, kidney, and heart. Mice appear to be more
sensitive to selenium with regard to lung toxicity than rats. Rats may be more
sensitive to the cardiotoxic effects, with an LEL of 0.1 mg/kg/d in a chronic study
(the study had some deficits in study design) (ATSDR, 1989).
Hematological effects have been observed in multiple acute and chronic animal
studies No human studies were located for this report or by ATSDR in their
literature review. Rats subchronically exposed to wheat containing selenium at a
dose of 0.68 mg/kg/d for 6 weeks had a reduction of blood hemoglobin. At 0.75
mg/kg/d in a similar study, red cell hemolysis was observed (ATSDR, 1989).
Bone softening in livestock has been noted with an LEL of 0.2 mg/kg/d with
exposure over several months (less than 100 days). Adverse effects on the liver
have been observed in multiple animal studies with LELs of 0.8 mg/kg/d and
above Kidney damage has also been noted with an LEL of 0.31 mg/kg/d. Dermal
effects have been observed at doses as low as 0.053 mg/kg/d in humans with
dietary exposure (ATSDR, 1989). This observation served as a partial basis for the
calculation of an MRL by ATSDR. Depression of the immune system was
observed in rats exposed subchronically to sodium selenite at 0.75 mg/kg/d. At
lower doses (0.075 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. No NOEL was identified in the study (ATSDR, 1989).
Chronic exposure studies in animals have identified multiple adverse effects on the
reproductive ability of animals and on offspring viability. Effects include: reduced
rates .of conception at 0.41 mg 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 0.34 mg/kg/d, increased fetal resorption
and decreased conception rate in livestock exposed at an LEL of approximately
0 5 mg/kg/d, failure to breed in a three-generation study of mice exposed at 0.42
mg/kg/d no effects in a two-generation study of mice at 0.21 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 did not appear to be affected in
the results reported, although the testes are a storage site for selenium (ATSDR,
1989).
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Neurological symptoms have been reported in human and animal studies. A family
exposed to approximately 0.26 mg/kg/d via drinking water reported various
symptoms of selenosis including listlessness and a lack of mental alertness.
Effects ceased when the water use was discontinued. More severe effects have
been observed in high-selenium areas of China. Peripheral anesthesia and pain
in the limbs were reported, although no associated estimate of exposure was
provided. Exaggerated tendon reflexes, convulsions, paralysis, and hemiplegia
were estimated to occur at a minimum chronic exposure of 0.053 and an average
of 0.083 mg/kg/d. A NOAEL of 0.025 was estimated. This information was used
by ATSDR to calculate a chronic exposure MRL of 0.003 mg/kg/d (ATSDR, 1989).
Neurological effects identified in animal studies include: drowsiness, lethargy,
ataxia, paralysis, bilateral lesions in the spinal cord, impaired vision, aimless
wandering behavior, and neuronal degeneration of the cerebral and cerebellar
cortices. Many of these were observed at relatively high doses; however, the
neuronal degeneration was observed at an LEL of 0.6 mg/kg/d dosing with sodium
selenite mixed in food (ATSDR, 1989).
The IRIS RfD and ATSDR MRL are within 1 order of magnitude of each other. The
IRIS value was used to calculate fish consumption limits shown in Section 4 for
chronic exposure toxicity. Please see the note at the end of the Developmental
Toxicity section for cautions regarding this use of these values.
5.2.4.5 Developmental Toxicity—
Limited information is available on the developmental toxicity of selenium in
humans. One anecdotal report indicated that selenium exposure may be
associated with spontaneous abortion and skeletal abnormalities (ATSDR, 1989);
however, the anecdotal nature of the report makes it inappropriate for drawing
conclusions regarding causality.
In animals, selenium has caused growth retardation, decreased fertility, embryo-
toxicity, fetotoxicity, and teratogenic effects. One researcher noted that, in a high-
selenium area, teratogenic effects were not seen in humans, but they were
observed in chickens (IRIS, 1993).
A multigeneration 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 LOEL
can be obtained from this study. A one-generation mouse study found a NOEL of
0.39 mg/kg/d. An early five-generation study identified a NOEL of 0.075 mg/kg/d
and a LOEL of 0.125 mg/kg/d with a 50 percent reduction in the number of young
reared at that dose. There are multiple possible reasons for the reduction,
including decreased fertility; consequently, it is not appropriate for use in cal-
culating an exposure limit for developmental effects. A recent study in primates
identified no developmental effects up to 0.3 mg/kg/d. However, the study utilized
dosing over a portion of the pregnancy, and, unlike the multigenerational studies,
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it did not include dosing prior to and during all of the pregnancy or dosing of the
neonates (IRIS, 1993).
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, 1989). At 0.41 mg, pigs exposed
from 8 weeks of age had offspring with significantly reduced birth weight and
weaning weights (ATSDR, 1989).
ATSDR has reported studies on experimental animals that have yielded the
following results: prenatal exposure at 0.34 mg/kg/d caused reduced fetal growth
with a NOAEL of 0.17 mg/kg/d; mice exposed to 0.42 mg/kg/d for three
generations had an increased incidence in fetal deaths and a high proportion of
runts among survivors; macaques exposed prenatally at levels up to 0.3 mg/kg/d
exhibited no adverse effects. It was noted that exposure to inorganic selenium
compounds at levels that are not maternally toxic have not produced teratogenic
effects (ATSDR, 1989). (EPA's guidelines on developmental toxicity specify that
dosing should include doses that cause some level of maternal toxicity; therefore,
this is not cause for dismissing the study results.)
Based on the reviewed information, the multigeneration mouse study cited in IRIS
with a LOEL of 0.39 mg/kg/d'appears to be the most appropriate value for
calculating an estimated exposure limit for developmental effects because there
are no other appropriate studies that provide data on long-term maternal and
offspring exposure effects. There is concern regarding the use of these results
because severe effects were seen at the LOEL and because severe effects have
been observed in other studies at approximately the same exposure level. The
standard uncertainty factors used to calculate an estimated exposure limit would
typically take into consideration inter- and intraspecies variability and the use of a
LOEL rather than a NOEL. A modifying factor for the severity of effects at the
LOEL could also be applied. The resulting value is within 1 order of magnitude of
an exposure limit that could be calculated from the NOEL of 0.17 mg/kg/d for
reduced fetal growth (as reported by ATSDR). Due to the longer-term nature of the
dosing, the multigeneration study cited in IRIS may be more appropriate.
Note: Decisions regarding thresholds for adverse effects of selenium are complex
because selenium is an essential nutrient. Consequently, the application of
uncertainty factors in the standard manner may not be appropriate. Some
exposure to selenium is necessary, as indicated by the RDA. There appears to be
a relatively small margin between the effective/necessary dose and the toxic dose
for this chemical. Additionally, the need for selenium and the toxicity of selenium
is expected to vary among individuals. Consequently, it is necessary to evaluate
the overall exposure to selenium in order to evaluate potential risks and make well-
informed decisions regarding exposure limits. Decisions regarding the contribution
to total selenium exposure that can come from fish without generating toxicity will
depend on the cumulative exposures from other sources. This is expected to vary
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considerably depending on the part of the country in which individuals reside, their
dietary habits, and other factors. If these factors were not a consideration for
selenjum exposure, an additional modifying factor would be recommended when
estimating exposure limits for developmental effects due to the serious nature of
effects observed in multiple species at or near the LEL of 0.39. Readers should
carefully review the toxicity data regarding selenium and determine the appropriate
exposure limit for developmental effects, based on the exposures anticipated in
their States and their interpretation of the toxicological and epidemiological
literature. See also Abernathy et al. (1993) for additional guidance on this topic.
It will be necessary to obtain a NOEL from a multigeneration study and to further
explore the mechanisms of fetal and neonatal lethality associated with selenium
exposure to adequately determine the appropriate exposure limit for develop-
mental effects. A well-designed human epidemiological study of prenatally
exposed individuals from high naturally occurring selenium areas is needed to
provide insight into human effects of selenium exposure.
5.2.4.6 Mutagenicity—
There are many positive mutagenicity assays on selenium compounds including
unscheduled DNA 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, 1993).
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, 1989). For a summary of study results, see the Toxicological Profile for
Selenium (ATSDR, 1989).
5.2.4.7 Carcinogenicity—
EPA has determined that there are insufficient data to assess the carcinogenic
potency of selenium. EPA has classified selenium sulfide as a probable human
carcinogen (B2), based on liver and lung tumors in oral exposure studies in
multiple species (IRIS, 1993). Some human studies indicated that combined
vitamin E and selenium deficiencies may lead to higher cancer risks (ATSDR
1989).
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 popu-
lations with elevated exposures arising from exposure via food produced in high-
selenium areas (ATSDR, 1989).
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Based on the occurrence of adverse effects reported in human and animal studies,
individuals with diseases or disorders of the following organ systems may be at
greater risk from selenium exposure than the general population: hematopoietic,
dermal, nervous, liver, kidney, cardiac, and immune systems.
HSDB listed individuals with the following conditions as requiring additional
protection: chronic indigestion or a history of peptic ulceration; skin, lung, kidney
or liver disease; dermatitis; chronic bronchitis; skin allergy or respiratory tract
infection; jaundice; or albuminuria (HSDB.1993). Their cautions are based on all
routes of selenium exposure.
5.2.4.9 Interactive Effects-
Selenium alters the toxicity of many chemicals. It reduces the toxicity of mercury,
cadmium, lead, silver, and copper; some forms reduce arsenic toxicity. Detailed
information on specific interactions can be found in the Toxicological Profile for
Selenium (ATSDR, 1989). Selenium also interacts with vitamins, sulfur-containing
amino acids, xenobiotics, and essential and nonessential elements. ATSDR notes
that most interactions are beneficial (ATSDR, 1989).
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, developmental studies
regarding cataract formation, immunotoxicity, neurotoxicity, especially behavioral
and histopathological CMS effects, pharmacokinetic, and bioaccumulation, and
bioavailability from environmental media (ATSDR, 1989). A multigeneration study
that utilizes sensitive endpoints for toxicity is needed to develop a more adequately
based exposure limit for developmental effects.
5.2.4.11 Summary of EPA Levels of Concern-
Chronic Toxicity
Carcinogenicity
5x10"3mg/kg/d
Insufficient data to assess carcinogenicity. Note that selenium
sulfide is classified as a Group B2 carcinogen.
5.2.4.12 Major Sources—
ATSDR (1989), HSDB (1993), IRIS (1993).
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
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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, 1986p. 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, 1986f). 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
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,
1992e).
5.2.5.3 Acute Toxicity—
The limited available data show tributyltin to be quite toxic to animals, with LD50s
ranging between 122 and 194 mg/kg in rats (ATSDR, 1992; HSDB, 1997). No
other information is available on the acute effects of tributyltin oxide.
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, 1997).
IRIS provides an RfD for tributyltin oxide of 3.0 x 10'5 mg/kg/d, based on a NOAEL
of 0.025 mg/kg/d and an uncertainty factor of 1,000. This was based on a chronic
rat feeding study in which immunotoxicity was observed. The uncertainty factor of
1,000 reflects the uncertainty in extrapolating from laboratory animals to humans,
the uncertainty in the range of human sensitivity, and the uncertainty due to the
lack of important toxicological data (IRIS, 1997).
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EPA has medium confidence in the studies on which the RfD was based, low
confidence in the database, and low confidence in the RfD. This is based on the
fact that the principal study was a well-designed and well-conducted chronic
toxicity assay; however, the number of animals used in the study was somewhat
minimal and only a preliminary report of the study was available for review. The
low ranking in the database is due to a lack of independent confirmation of the
critical effect, the lack of toxicological data for a second species, and the lack of
information on reproductive toxicity (IRIS, 1997).
5.2.5.5 Developmental Toxicity—
No studies are available on the developmental effects of tributyltin oxide in
humans. A study in mice reported dose-related decreases in fetal weights and
some skeletal abnormalities, such as fused ribs and cleft palates, at all dose levels
and also in the controls (ATSDR, 1992). When pregnant rats were exposed to high
doses of tributyltin oxide, decreased numbers of live births and decreased growth
and viability of the offspring were reported (HSDB, 1997).
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, 1992e; HSDB, 1997).
5.2.5.7 Carcinogenicity—
There are very limited data on the carcinogenicity of tributyltin oxide. No human
studies are available and the one available animal study noted an increased
incidence of some benign tumors at the highest dose level in rats. The authors
concluded that their results could not be considered evidence of carcinogenicity,
but that the changes may be related to a direct action of tributyltin oxide on the
endocrine glands (ATSDR, 1992; HSDB, 1997). EPA has not classified tributyltin
oxide for carcinogenicity.
5.2.5.8 Special Susceptibilities—
ATSDR reported that no studies were located regarding unusual susceptibility of
any human subpopulation to tributyltin oxide. However, based on the target organ
systems of organotin compounds, persons with liver disease, blood disorders,
deficiencies of the immune system, neurobehavioral disorders, and perhaps
kidney disease could be predisposed to adverse health effects of tributyltin oxide
under appropriate conditions of exposure (ATSDR, 1992).
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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).
5.2.5.10 Critical Data Gaps—
The following are areas where data gaps exist for tributyltin oxide: acute,
intermediate (14 to 365 days), and chronic exposures and reproductive,
developmental, and neurotoxic studies.
5.2.5.11 Summary of EPA Levels of Concern—
Chronic Toxicity 3.0 x 10"5 mg/kg/d
Carcinogenicity Insufficient data to determine carcinogenic status.
5.2.5.12 Major Sources—
ATSDR (1992), HSDB (1997), IRIS (1997), U.S. EPA (1986f).
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5.3 ORGANOCHLORINE PESTICIDES
5.3 ORGANOCHLORINE PESTICIDES
In addition to the discussions of individual target analytes, please refer to the
discussion of toxicity characteristics of the organochlorine group in Appendix C.
5.3.1 Chlordane
5.3.1.1 Background—
Chlordane is an .organochlorine pesticide comprised of the sum of cis- and trans-
chlordane and trans-nonachlor and oxychlordane for purposes of health advisory
development (U.S. EPA, 1993a). It was used extensively until most uses were
banned in 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 bioaccumulates in biological materials (IRIS, 1993). It is highly lipophilic
and readily absorbed via all routes. 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 (ATSPR, 1992d).
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, 1992d).
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, 1992d).
5.3.1.3 Acute Toxicity—
Chlordane is moderately to highly toxic with an estimated lethal dose to humans
of 6 to 60 g (IRIS, 1993). See the listing of usual effects associated with organo-
chlorine exposure in Appendix C.
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5.3.1.4 Chronic Toxicity—
Chlordane has classic organochlorine toxicity as described in Appendix C. The
principal systems affected by exposure are liver, nervous system, and immune
system. Other effects include neurological abnormalities including grand mal
seizures and altered EEG results (ATSDR, 1992d).
Reduced fertility and survivability in mice and rats has occurred at 25 and 16
mg/kg, respectively, and may be associated with reduced binding of progesterone
in the endometrium or with altered metabolism and circulating levels of steroid
hormones. The studies were not designed to identify thresholds or mechanisms
for action (ATSDR, 1992d) and cannot be used to derive dose-response data for
estimation of an RfD.
Jaundice has been reported in humans living in homes treated with chlordane for
termite control. Chemistry changes indicative of altered liver function were
observed in pesticide applicators in Japan who were exposed to chlordane
(ATSDR, 1992d).
Multiple neurological effects have been reported in humans exposed both acutely
and chronically. According to ATSDR, neither animal nor human studies have
evaluated subtle neurological or behavioral effects that may occur at low levels.
Consequently, it is not possible to assess the likelihood of human effects at
environmental exposure levels (ATSDR, 1992d).
IRIS provides an RfD of 6.0 x 10"5 based on a NOAEL of 0.055 mg/kg/d in a study
that found liver atrophy in female rats. The standard uncertainty factors of 10 each
for inter- and intraspecies variability were applied. An additional safety factor of 10
was applied "to account for the lack of an adequate reproduction study and
adequate chronic study in a second mammalian species, and the generally
inadequate sensitive endpoints studied in the existing studies, particularly since
chlordane is known to bioaccumulate over a chronic duration" (IRIS, 1993).
Confidence in this RfD is low for these reasons (IRIS, 1993).
5.3.1.5 Developmental Toxicity—
According to the IRIS file, "there have been 11 case reports of CNS effects, blood
dyscrasias and neuroblastomas in children with pre/postnatal exposure to
chlordane and heptachlor" (IRIS, 1993). Data were insufficient to calculate an
exposure limit for developmental effects from this study.
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
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corticosterone levels were observed, which may result from a change in the
neuroendocrinological feedback mechanisms (ATSDR, 1992d).
There is insufficient information to develop a well-based estimated exposure limit
for developmental effects. According to ATSDR, neither animal nor human studies
have evaluated subtle neurological or behavioral effects that may occur at low
levels. Consequently, it is not possible to assess the likelihood of human effects
at environmental exposure levels (ATSDR, 1992d). Neurological and behavioral
effects may be the most sensitive measures of chlordane developmental toxicity.
This appears to be the case for some other organochlorine pesticides (see DDT
and toxaphene). However, it is not possible to estimate the threshold level
because the LOEL caused multiple and serious effects. If readers elect to calculate
an exposure limit for developmental effects, it should be considered a limited
estimate due to the lack of information on the threshold for effects. The standard
uncertainty factors used in this calculation would typically take into consideration
inter- and intraspecies variability, the use of a LOEL rather than a NOAEL, and the
poor quality of the database.
Chlordane accumulates in body tissues; 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 chlordane, even if exposure is reduced
during pregnancy, the outcome of that pregnancy may be affected, depending on
the timing and extent of prior exposure.
Regarding cancer in children, see the discussion in Section 5.3.1.7.
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, 1993).
5.3.1.7 Carcinogenicity—
Chlordane is classified as a probable human carcinogen (B2) by EPA based on
oral studies in animals. The oral cancer slope factor of 1.3 per mg/kg/d is the
geometric mean of the cancer potencies calculated from four data sets (IRIS,
1992). This value was used to develop the fish consumption limits for carcinogenic
toxicity listed in Section 4.
Positive results have been obtained in four strains of mice of both sexes and in
male rats. In addition, numerous structurally related organochlorine pesticides
have been found to be carcinogenic.
Neuroblastoma and acute leukemia have also been associated with prenatal and
early childhood exposure to chlordane (ATSDR, 1993c).
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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, 1992d).
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, 1992d).
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. In addition, it has been hypothesized that
a subpopulation may exist with a predisposition to blood dyscrasias resulting from
chlordane exposure. Identification of such a population is not now possible
(ATSDR, 1992).
5.3.1.9 Interactive Effects—
Chlordane is a potent inducer of hepatic microsomal enzymes. (See a discussion
of organochlorine effects related to this induction in Appendix C.) 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, 1992d).
It has been suggested that increased dietary vitamins C or E or selenium may be
protective against free-radical-induced toxicity (ATSDR, 1992d).
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).
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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, 1993).
It is clear from this list that the developmental effects of chlordane have not been
adequately evaluated.
According to ATSDR, neither animal nor human studies have evaluated subtle
neurological or behavioral effects that may occur at low levels. These types of
studies are needed to assess the likelihood of human effects at environmental
exposure levels (ATSDR, 1992d).
ATSDR has declined to develop oral MRLs for acute, intermediate, or chronic
duration oral exposure due to the lack of data on sensitive endpoints for these
durations. They note the need for a behavioral study because it appears to be a
sensitive endpoint. 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, 1992d).
5.3.1.11 Summary of EPA Levels of Concern-
Chronic Toxicity
Carcinogenicity
6x1(r5mg/kg/d
1.3 per mg/kg/d.
5.3.1.12 Major Sources—
ATSDR (1992d), HSDB (1993), IRIS (1993).
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 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- year-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,
1992c), the sum of the 4,4'- and 2,4'- isomers of DDT, DDE, and ODD should be
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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—
See the listing of usual effects associated with organochlorine exposure in
Appendix C. 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).
DDD is considered less toxic than DDT in animals. Symptoms develop more slowly
and have a longer duration with DDD than with DDT exposure. Lethargy is more
significant and convulsions are less common than with DDT exposure (HSDB,
1993).
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 (categorized as serious
effects by ATSDR) at 0.18 mg/kg/d. Other effects were observed at higher doses.
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No studies were provided on immunological effects following chronic exposure
(ATSDR, 1992c).
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. NOELs are not available for these studies.
Other studies have identified NOELs ranging from 2.4 to 10 mg/kg/d with severe
effects at 12 mg/kg/d (increased maternal and offspring death) (ATSDR, 1992c).
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).
IRIS lists an oral RfD of 5 x 10"4 mg/kg/d based on liver effects with a NOEL of
0.05 mg/kg/d from a 27-week 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, 1993). The corroborating study
was conducted in 1948.
More recent studies of the immunological and reproductive systems (noted above)
suggest a LOEL from subchronic studies in the range of 0.18 to 0.35. There are
numerous studies supporting the occurrence of both types of effects, and both are
serious in nature. An alternative estimated exposure limit could be calculated using
these more recent data. The most sensitive endpoint appears to be immunological
effects observed in the rabbit study (noted above). This study had a LOEL of 0.18
mg/kg/d. The standard uncertainty factors used in this calculation would typically
take into consideration inter- and intraspecies variability, the use of a LOEL rather
than a NOAEL, and the use of a less-than-lifetime study.
5.3.2.5 Developmental Toxicity—
DDT causes embryotoxicity and fetotoxicity but not teratogenicity in experimental
animals (ATSDR, 1992c). Studies indicate that estrogen-like effects on the
developing reproductive system occur (ATSDR, 1992c). 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,
1992c). 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).
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A three-generation rat reproduction study found increased offspring mortality at all
dose levels with a LOEL of 0.2 mg/kg/d. Three other reproduction studies found
no effects at much higher dose levels (IRIS, 1993). 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 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, placenta! 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). It is not known whether this is related to the reproductive system toxicity of
DDT (see Section 5.3.2.4) or developmental toxicity. 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.
ATSDR reports that a recent developmental study in mice found behavioral
abnormalities in offspring exposed prenatally at 0.5 mg/kg/d. Latent effects were
observed following cessation of exposure, and subsequent tissue evaluation found
structural/function alterations in the brain. Effects reported include an abnormal
increase in activity and probable altered learning ability. The effects occurred at
levels approximately 50-fold lower than those that were noted in adults and did not
cease when dosing was discontinued or when tissue levels had decreased. This
information was used to support the hypothesis of permanent structural changes
in the brain. The results of this study were used by ATSDR to calculate an acute
exposure MRL of 5 x 10"4 mg/kg/d using standard uncertainty factors of 10 each
for inter- and intraspecies variability and the use of a LOEL rather than a NOEL
(ATSDR, 1992c). This MRL is based upon a sensitive endpoint with structural and
functional toxicity correlations and should be considered for use as an exposure
limit for developmental effects of DDT, DDE, and DDD. Readers may elect to
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consider the ATSDR MRL for developmental toxicity. The MRL is the same value
as the current IRIS RfD (as listed under Section 5.3.2.4).
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.
5.3.2.6 Mutagenicity—
"Genotoxicity studies in human systems strongly suggest that DDT may cause
chromosomal damage" (ATSDR, 1992c). This is supported by in vitro and in vivo
studies in animals (ATSDR, 1992c) and in some bacterial assays (HSDB, 1993).
There are multiple positive assays including human lymphocytes, human
leukocytes, human fibroblasts, an oncogenic transformation, and unscheduled
DNA synthesis in rats in multiple studies (ATSDR 1992c; 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, 1993). 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, 1993). 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
Carcinogenicity 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
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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) (see also Appendix C). 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 drugs' efficacy and/or timing if they
are exposed 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, 1992c). Based on
information discussed above regarding 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. (See the discussion of
organochlorine effects related to induction of the mixed function oxidase system
in Appendix C.) DDT is reported to promote some tumorigenic agents and
antagonize others. The actions may be related to the induction of microsomal
enzymes (ATSDR, 1992c).
5.3.2.10 Critical Data Gaps—
IRIS notes the lack of a NOEL for reproductive effects and a relatively short
duration for the critical study on which the RfD is based. No intermediate or chronic
oral MRLs were calculated by ATSDR because of the lack of a NOEL and the
seriousness of the LOEL in significant studies (ATSDR, 1992c).
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
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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, 1992c).
5.3.2.11 Summary of EPA Levels of Concern—
These values should be used for the sum of the 4,4'- and 2,4'- isomers of DDT,
DDE, and ODD.
Chronic Toxicity 5 x 10'4 mg/kg/d
Carcinogenicity 0.34 per mg/kg/d.
5.3.2.12 Major Sources—
ATSDR (1992c), Hayes (1982), HSDB (1993), IRIS (1993).
5.3.3 Dicofol (Kelthane)
5.3.3.1 Background—
Dicofol is an organochlorine pesticide that is structurally similar to DDT and is
frequently contaminated with isomers of DDT, DDE, and ODD (U.S. EPA, 1993a).
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—
Very few data were located regarding the pharmacokinetics of dicofol. Due to its
structural similarity to DDT, it may be assumed to have some of the same
properties. Data regarding metabolites are not consistent. The mechanism of
action is hypothesized to be inhibition of the ATPase associated with oxidative
phosphorylation and cation transport in the plasma membranes (HSDB, 1993).
5.3.3.3 Acute Toxicity—
See the listing of usual effects associated with organochlorine exposure in
Appendix C. The acute oral LD50s for dicofol from animal studies ranged from 640
to 1,810 mg/kg (U.S. EPA, 1993g).
5.3.3.4 Chronic Toxicity—
No IRIS file was located for this chemical. OPP lists an RfD of 0.001 mg/kg/d
based on a NOEL of 1 mg/kg/d in a 2-year rat feeding study (no information was
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located on the critical effect). Uncertainly factors totaling 1,000 were applied (U.S.
EPA, 1992d).
The liver is a target organ for dicofol both for systemic and carcinogenic effects.
Studies have also reported thyroid hypertrophy in rats at 25 mg/kg/d. A NOEL of
0.9 mg/kg/d was identified in a recent study of liver toxicity, based on gross and
microscopic pathology and enzyme alterations, in a 1-year dog study (U.S. EPA,
1993g). This study would yield an estimated exposure limit within approximately
1 order of magnitude of the RfD listed above.
Due to the limited information available for this review on the dose-response
dynamics for dicofol, it is recommended that the OPP value of 0.001 mg/kg/d be
used for chronic systemic toxicity.
5.3.3.5 Developmental Toxicity—
Two three-generation reproductive studies in mice and rats both identified a NOEL
of 1.5 mg/kg/d with effects at 3.375 mg/kg/d noted as reduced litter size, reduced
body weight, and reduced offspring survival (U.S. EPA, 1993g). The reviewed data
did not contain information regarding underlying mechanisms of fetal or neonatal
toxicity. Additional uncertainty arises because of the limited information available
in the database regarding the study outcomes. They are gross measures of toxicity
and do not provide any indication of the level of exposure at which organ toxicity
that led to death was occurring. Consequently, an estimated exposure limit for
developmental effects cannot be estimated with precision. If these studies were
used, the standard uncertainty factors employed in the calculation would typically
take into account consideration of inter- and intraspecies variability. An additional
modifying factor for the limited information available in the database could also be
used.
As with the other organochlorines, it is anticipated that dicofol 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.
5.3.3.6 Mutagenicity—
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).
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5.3.3.7 Carcinogenicity-
Dicofol has been classified as a B2 and C carcinogen by offices within EPA. OPP
lists the potency value as 0.44 per mg/kg/d (U.S. EPA, 1992c). The EPA
Carcinogenicity Assessment Group (CAG) has recommended that 0.34 per
mg/kg/d be used for combinations of dicofol with DDT, DDE, and ODD (U.S. EPA,
1993a). The value of 0.44 per mg/kg/d was used to develop fish consumption limits
listed in Section 4 for carcinogenic effects.
5.3.3.8 Special Susceptibilities-
Individuals taking medications that involve the mixed function oxidase system may
need to alter their dosages when exposure to dicofol is occurring at significant
levels. No specific information was available on the critical dosage for interaction.
See Appendix C for more information on this topic.
Individuals with liver disease and children exposed prenatally may also be at risk
based on the toxicity information reviewed.
5.3.3.9 Interactive Effects—
As with other organochlorine pesticides, microsomal enzyme induction occurs and
may cause interactions with other chemicals. See a discussion of this in Appendix
C. No additional data were located.
5.3.3.10 Critical Data Gaps-
Information is lacking on neurotoxicity endpoints for chronic and developmental
toxicity. Based on data available on other organochlorines, this type of toxicity
commonly occurs and may be a sensitive endpoint that could serve as a useful
basis for chronic or developmental toxicity exposure limits. The reviewed data did
not contain information regarding underlying mechanisms of fetal lethality. A
sensitive measure of developmental toxicity is necessary to generate a protective
exposure limit. Clarification is also needed regarding the carcinogenic nature of
dicofol.
5.3.3.11 Summary of EPA Levels of Concern-
Chronic Toxicity
Carcinogenicity
1.0 x10'3 mg/kg/d
0.44 per mg/kg/d for dicofol alone
0.34 per mg/kg/d in combination with DDT, DDE, DDD.
5.3.3.12 Major Sources—
HSDB (1993), U.S. EPA (1993g).
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5.3.4 Dieldrin
5.3.4.1 Background—
Dieldrin is an organochlorine pesticide that was phased out between 1974 and
1987. It continues to be detected nationwide due to its relatively long half-life.
Dieldrin is also a product of aldrin metabolism (ATSDR, 1991 a).
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 a).
In human dosing studies 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 a).
i
The metabolism of dieldrin is described in detail in ATSDR (1991 a). 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).
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 a).
5.3.4.3 Acute Toxicity—
See the listing of usual effects associated with organochlorine exposure in
Appendix C. Additional effects include: possible hematological effects in humans
(pancytopenia and thrombocytopenia, immunohemolytic anemia) (ATSDR, 1991 a).
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. Uncertainty factors
of 10 each for inter- and intraspecies variability were applied (IRIS, 1993). Liver
toxicity has been observed in multiple animal studies and in human acute
exposure episodes. Adaptive changes (e.g., liver, enlargement) have been
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observed at 0.00035 mg/kg/d in a subchronic rat study. ATSDR has calculated an
MRLthat is equal to the RfD listed in IRIS (ATSDR, 1991 a).
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, 1993). 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 a). 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 CNS
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 a).
Neurotoxicity has been observed in humans with chronic inhalation and dermal
exposures (ATSDR, 1991 a). Chronic exposure of pesticide applicators to dieldrin
led to idiopathic epilepsy, which ceased when exposure was terminated (HSDB,
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 CNS (as measured by EEG), peripheral nerve
activity, or muscle activity (ATSDR, 1991a).
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 a). Neural lesions (cerebral, cerebellar, brainstem,
and vascular) were observed in chronically exposed rats at 0.004 mg/kg/d (HSDB,
1993).
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With the exception of the neurological study discussed directly above, the
information reviewed regarding neurotoxicity indicates that the IRIS RfD would be
protective against adverse effects, using standard assumptions for the
development of an exposure limit. Although the neurological rat study cited above
noted effects at 0.004 mg/kg/d, the human study of exposure over an 18-month
period at 0.003 mg/kg/d found no effects on the CNS based on various sensitive
measures. Taking the results of the human study under consideration, it appears,
based on the information reviewed, that the IRIS RfD provides adequate protection
against neurological effects in the human population.
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 a). See also notes regarding estrogenic
activity in Section 5.3.4.7.
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 an LEL 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 a).
Based on the information reviewed regarding reproductive toxicity, it appears that
the IRIS RfD would be protective against adverse effects, using standard
assumptions and uncertainty factors for calculating an estimated exposure limit.
5.3.4.5 Developmental Toxicity—
IRIS provides limited information regarding the developmental toxicity of dieldrin.
A NOEL 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 LEL 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 LEL
dose (IRIS, 1993). An RfD based on developmental effects is not provided in IRIS.
A variety of effects in multiple organ systems have been observed in experimental
animals exposed prenatally to dieldrin. Skeletal anomalies and malformations
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(e.g., cleft palate, webbed foot; open eyes, extra ribs) were identified at relatively
large doses (LEL of 3 mg/kg/d) (ATSDR, 1991 a).
Abnormalities of the CNS, eye, and ear were noted with a TD Lo (similar to a
LOEL) 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 a). Note that liver lesions are the basis for the chronic toxicity RfD
derived from a study of adult animals, as reported in IRIS (IRIS, 1993). A
multigeneration study in mice found histological changes in liver, kidney, lungs,
and brain tissues in the first and second generation offspring at an LEL 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 a).
Neural lesions in prenatally exposed rats were found at an LEL 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 a). In a rat multigeneration study, a TD Lo 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
r 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 LEL of 0.004 mg/kg/d provides the
most sensitive developmental toxicity measure of those reviewed. If the LEL 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 LEL rather than a NOAEL.
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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.
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 DNA; however, it may act by depressing transfer
RNA activity, increasing unscheduled DNA 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 a).
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) were used to develop the cancer potency (IRIS, 1992). This
value was used to calculate fish consumption limits listed in Section 4 for
carcinogenic effects.
A variety of tumor types have been observed in animal studies including
pulmonary, lymphoid, thyroid, and adrenal (ATSDR, 1991 a). ATSDR has
concluded that dieldrin is probably a tumor promotor, based on genotoxicity and
mechanistic studies reviewed (ATSDR, 1991 a). Dieldrin 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 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).
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5.3 ORGANOCHLORINE PESTICIDES
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 a). 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.
See also a discussion of susceptibilities associated with pharmaceutical use in
Appendix C.
5.3.4.9 Interactive Effects-
See the discussion of organochlorine effects related to induction of the mixed
function oxidase system in Appendix C. 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: 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 a).
5.3.4.11 Summary of EPA Levels of Concern-
Chronic Toxicity
Carcinogenicity
5x10"5mg/kg/d
16permg/kg/d.
5.3.4.12 Major Sources—
ATSDR (1991 a), HSDB (1993), IRIS (1993).
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5.3 ORGANOCHLORJNE PESTfCfDES
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 found widely in food samples, including one
of 10 fruit and fruit juice samples for infants at a mean concentration of 0.01 ppb
(HSDB, 1993).
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, 1993b).
5.3.5.3 Acute Toxicity—
Endosulfan has a high acute toxicity to humans, with an estimated lethal dose of
50 to 500 mg/kg. Multiple animal studies found females much more sensitive to
exposure than males (e.g., acute oral LD50of 9.5 in females and 40.4 in males)
(U.S. EPA, 1992c). See the listing of usual effects associated with organochlorine
exposure in Appendix C. In addition to those listed in Appendix C, bluing of the
skin (IRIS, 1993), hematopoietic system damage and anemia, possibly damage
to red blood cell membranes, cardiac toxicity, and immunotoxicity have been noted
(ATSDR, 1993b).
5.3.5.4 Chronic Toxicity—
IRIS previously provided an RfD of 5 x 1Q-5 mg/kg/d for endosulfan based on a
LOAEL of 0.15 mg/kg/d from a two-generation rat reproduction study that identified
kidney toxicity. Uncertainty factors totaling 3,000 were applied (IRIS, 1992). The
RfD was withdrawn in December 1992, and a new RfD summary is under
development (IRIS, 1993). The Office of Pesticide Programs has recently
reevaluated this chemical and calculated an RfD of 6 x 10"3mq/kq/d (U S EPA
1996b).
ATSDR developed intermediate exposure duration (14-365 days) and chronic
duration MRLs of 0.002 mg/kg/d for both intermediate and chronic exposures.
These MRLs are based on immunotoxicity and hepatotoxicity, respectively
(ATSDR, 1993b). y
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5.3 ORGANOCHLORINE PESTICIDES
Other chronic effects of enclosulfan noted in studies include: blood vessel
aneurysms at 0.65 mg/kg/d, neurological effects at 1.71 mg/kg/d, damage to the
hematopoietic system at 3.75 mg/kg/d, and elevated hemoglobin levels at 0.1
mg/kg/d (U.S. EPA, 1993i). It appears that the old IRIS RfD would be protective
against the effects noted, based on current risk assessment methods.
Two National Cancer Institute studies have identified the following effects:
interstitial fibrosis or acute tubular necrosis of the kidney, atrophy of the testes,
polyarteritis, parathyroid hyperplasia, osteitis fibrosis of the bone, and abscesses
of the lung. The kidney effects led to most deaths (dosages were not listed in the
database) (HSDB, 1993). A number of additional studies have also found damage
to the male reproductive system associated with exposure to endosulfan (e.g.,
testicular necrosis, aspermatogenesis, degeneration of seminiferous tubule
epithelium) (ATSDR, 1993b). (See discussion of estrogenic activity under
Carcinogenicity below.)
A neurological study in rats exposed at 3 mg/kg/d for 30 days found increased
aggressive behavior at both doses along with a significant increase in serotonin
binding in the frontal cortical membranes that may have been due to an increase
in the affinity of the serotonin receptors (HSDB, 1993). This effect has negative
implications for human behavior. Abnormal increases in behavior in prenatally
exposed animals have also been noted for other organochlorine pesticides (see
Appendix C on organochlorines); however, this level of mechanistic detail has not
been located for other organochlorines in the reviews conducted for this document.
5.3.5.5 Developmental Toxicity--
Multiple teratogenic effects were associated with endosulfan exposure in a rat
developmental toxicity study including webbed forelimb, clubbed hind limbs,
hypoplastic aortic arch, edema and lordosis, increased incidence of small 4th and
unossified 5th sternebrae, and decreased pup size and weight. An increased
incidence of misaligned vertebrae was observed at all dose levels with an LEL of
0.66 mg/kg/d (U.S. EPA, 1993i)..
Other developmental studies have yielded a variety of results that are often
inconsistent (e.g., two separate studies found unspecified effects at the lowest
dose tested of 5 mg/kg/d in one study and no effects at the highest dose tested of
1.8 mg/kg/d in another study). A range-finding single-generation reproductive study
found increased liver weights at the lowest dose tested of 2.5 mg/kg/d. A two-
generation reproductive study found increased pituitary and uterine weights at 3.75
mg/kg/d and a NOEL of 0.75 mg/kg/d. Kidney discoloration, which was originally
attributed to hematopoietic damage at all doses, has been reevaluated and is now
considered by OPP to be part of the elimination process rather than an adverse
effect (U.S. EPA, 1993i). It is not clear why significant differences in effects were
noted in multiple recent rat studies (i.e., the first and last studies discussed in this
paragraph).
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5.3 OFtGANOCHLORlNE PESTICIDES
Postnatal exposure of rats for 5 weeks at 1 mg/kg/d resulted in aggressive
behavior and increased serotonin binding. This persisted after the dosing stopped.
The study authors concluded that the developing study subjects had a greater
sensitivity to endosulfan than adults (ATSDR, 1993b).
In the absence of more complete information, a conservative approach is
recommended for calculation of an estimated limit for developmental effects, due
to the severity of effects observed in the teratogenicity study with a LOEL of 0.66
mg/kg/d (listed first above). If this study, which appears to be the most sensitive,
were used to calculate an estimated exposure limit, the uncertainty factors used
in this calculation would typically take into consideration inter- and intraspecies
variability and the use of an LEL rather than a NOEL.
As with the other organochlorines, it is anticipated that endosulfan 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 endosulfan, even if exposure is reduced during pregnancy, the
outcome of that pregnancy may be affected, depending on the timing and extent
of prior exposure.
5.3.5.6 Mutagenicity—
Results of mutagenicity assays of endosulfan are mixed, with multiple positive and
negative studies (ATSDR, 1993b; HSDB, 1993; IRIS, 1993). 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, 1993b).
5.3.5.7 Carcinogenicity—
Insufficient information is available to determine the carcinogenic status of
endosulfan I and II. The carcinogenic assays have yielded mixed results, with
carcinomas, sarcomas, and lymphosarcomas identified at increased incidences
in some studies. ATSDR has concluded that the available animal study data were
negative or inconclusive (ATSDR, 1993b). Endosulfan 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 et al., 1993). In addition to potential carcinogenic effects,
endosulfan may also cause disruption of the endocrine system due to its
estrogenic activity (Soto et al., 1994).
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5.3 ORGANOCHLORINE PESTICIDES
5.3.5.8 Special Susceptibilities—
As noted above, multiple animal studies found females much more sensitive to
endosulfan exposure than males, some by nearly 1 order of magnitude (U.S. EPA,
1992c). However, toxicity studies indicate that the male reproductive system is a
target organ for endosulfan toxicity (ATSDR, 1993b).
Animal neurobehavioral study results indicate that the developing test animals had
a greater sensitivity to endosulfan than adults based on neurotransmitter patterns
and behavioral effects observed (ATSDR, 1993b). This may indicate that children
are at greater risk for neurotoxicity than adults. ATSDR has noted that:
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, 1993b)
Additional groups who may be at greater risk from endosulfan exposure 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, 1993b). See also a
discussion of susceptibilities associated with pharmaceutical use in Appendix C.
5.3.5.9 Interactive Effects-
Human anecdotal information suggests that endosulfan may act synergistically
with alcohol (ATSDR, 1993b). In laboratory animals, moderate protein deprivation
doubled the toxicity of endosulfan (Hayes and Laws, 1991).
Pentobarbital and endosulfan have demonstrated an interactive effect that is
probably related to microsomal enzyme activity. Endosulfan induces the mixed
function pxidase system (ATSDR, 1993b). Vitamin A inhibited the endosulfan-
induced activity of the mixed function oxidase system (ATSDR, 1993b). See a
discussion oforgaTibchloririeeffects "related to induction of the mixed function
oxidase system in Appendix C. '
5.3.5.10 Critical Data Gaps—
The increased susceptibility of females to endosulfan should be studied to
determine the underlying cause, evaluate whether the effect occurs with chronic
exposure, and identify a numerical modifier to adjust toxicity estimates and
exposure recommendations so that they provide adequate protection for females.
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; thus, they
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5.3 ORGANOCHLORJNE PESTICIDES
somewhat compromise the determination of an exposure limit for developmental
effects.
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, 1993b).
5.3.5.11 Summary of EPA Levels of Concern-
Chronic Toxicity
Carcinogenicity
5.3.5.12 Major Sources—
6 x10"3 mg/kg/d
Insufficient data to determine carcinogenic status.
ATSDR (1993b), HSDB (1993), IRIS (1993), U.S. EPA (1993J).
5.3.6 Endrin
5.3.6.1 Background-
Endrin is an organochlorine pesticide whose registration was canceled in 1984
(U.S. EPA, 1993a).
5.3.6.2 Pharmacokinetics—
Endrin, like the other organochlorine pesticides, is lipophilic. It bioaccumulates in
fat and probably brain tissue and can cross the placenta. Endrin is metabolized 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, 1990c).
5.3.6.3 Acute Toxicity—
Endrin has a high acute toxicity (IRIS, 1993). See the listing of usual effects
associated with organochlorine exposure in Appendix C. Blood pressure elevation
has also been noted (IRIS, 1993). The primary target of endrin is the central
nervous system (ATSDR, 1990c).
5.3.6.4 Chronic Toxicity—
IRIS provides an RfD of 2 x 1(T4 mg/kg/d based on a NOAEL of 0.025 mg/kg/d
from a 1969 chronic exposure dog study that identified histological lesions in the
liver and convulsions in study subjects exposed at the LEL of 0.05 mg/kg/d.
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5.3 ORGANOCHLORINE PESTICIDES
Uncertainty factors of 10 each for inter- and intraspecies variability were applied
(IRIS, 1993). ATSDR used the same study and safety factors to calculate an MRL
equal' to the IRIS RfD (ATSDR, 1990c).
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, 1993m). The LOAEL of
0.015 is within 1 order of magnitude of the LEL identified in the critical IRIS study.
The IRIS value was used to calculate fish consumption limits for chronic exposure
effects listed in Section 4.
5.3.6.5 Developmental Toxicity—
No developmental effects were listed in the IRIS file for endrin (IRIS, 1993).
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, 1990c). The variation in effects is
probably due in part to the different prenatal periods during which exposure
occurred (see ATSDR, 1990c). Reproductive outcome was adversely affected in
hamsters exposed to 1.5 mg/kg/d with decreased survival of pups (16 percent
mortality). The underlying cause was not discussed (ATSDR, 1990c).
Nervous system effects are a significant concern with organochlorine exposure.
In hamsters, abnormally increased pup activity in hamsters was observed with 1.5
mg/kg prenatal exposures for 9 days. The NOEL for these behavioral effects was
0.075 mg/kg/d (ATSDR, 1990c). In rats, increased activity was seen with prenatal
exposure to 0.3 mg/kg/d (ATSDR, 1990c). 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.
Both structural skeletal changes and neurological abnormalities are significant
developmental effects associated with endrin. Decreased survival, while a
significant effect, is not usually a sensitive measure of toxicity. The behavioral
effects observed with a NOEL of 0.075 mg/kg/d (discussed above) are
recommended for estimation of an exposure limit for developmental toxicity due
to their greater sensitivity. The uncertainty factors used in this calculation would
typically take into consideration inter- and intraspecies variability.
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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, 1990c).
5.3.6.7 Carcinogenicity—
Insufficient information is available to determine the carcinogenic status of endrin.
EPA has classified this as a Group D carcinogen (insufficient information
available). Some studies have yielded positive results and some studies that
reported negative results were considered to be inadequate (IRIS, 1993). Tumors
have been noted in the adrenal glands, pituitary glands, liver, mammary gland,
uterus, and thyroid in various studies and multiple species (IRIS, 1993). Endrin is
structurally related to a number of chemicals that are carcinogenic in test animals,
including chlordane, aldrin, dieldrin, heptachlor, and chlorendic acid (IRIS, 1993).
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, 1990c).
In addition, the skeletal and behavioral abnormalities associated with endrin
exposure in experimental animals indicate that prenatal exposure may generate
special risks.
Based on animal studies, females may be more susceptible than males to endrin-
induced toxicity (ATSDR, 1990c).
See also a discussion of susceptibilities associated with pharmaceutical use in
Appendix C.
5.3.6.9 Interactive Effects-
See a discussion of organochlorine effects related to induction of the mixed
function oxidase system in Appendix C. Dietary pretreatment with endrin
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5.3 ORGANOCHLORINE PESTICIDES
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; immunotox-
icity studies of animals and humans; human dosimetry studies; pharmacokinetic
studies; and studies of interspecies differences in metabolism and toxicity
(ATSDR, 1990c).
5.3.6.11 Summary of EPA Levels of Concern-
Chronic Toxicity
Carcinogenicity
3x10'4mg/kg/d
Insufficient data to determine carcinogenic status.
5.3.6.12 Major Sources—
ATSDR (1990c), IRIS (1993), U.S. EPA, 1993m.
5.3.7 Heptachlor Epoxide
5.3.7.1 Background—
Heptachlor epoxide is a breakdown product of the organochlorine pesticide
heptachlor and chlordane and is a contaminant of both products. It is more toxic
than either parent compound (ATSDR, 1993c). Although most uses of heptachlor
were suspended in 1978 and chlordane was removed from the market in 1988
(U.S. EPA, 1993J), 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, 1993c).
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, 1993c).
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5.3.7.3 Acute Toxicity—
See the listing of usual effects associated with organochlorine exposure in
Appendix C. The LD50s for heptachlor range from 40 to 162 mg/kg in rodents
(ATSDR, 1993C).
5.3.7.4 Chronic Toxicity—
IRIS provides an RfD of 1.3 x 1(r5mg/kg/d based on an LEL 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 LEL rather than a NOEL (IRIS, 1993). No additional
uncertainty factors were applied for the use of a less-than-lifetime study. The prin-
cipal study is of low quality and there is a low confidence in the RfD (IRIS, 1993).
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
1993c).
Significant changes in EEG patterns were found in female adult rats exposed to
1 and 5 mg/kg/d for three generations (ATSDR, 1993c).
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,
1993c). 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, 1993c).
5.3.7.5 Developmental Toxicity—
A 1973 two-generation dog reproductive study identified a NOEL of 0.025 mg/kg/d
with an LEL 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, 1993).
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
prior to mating and during gestation resulted in severely reduced pup survival (15
percent) at 21 days postpartum (ATSDR, 1993c). This is not a useful LOEL due
to the severity of effects observed at the lowest dose tested.
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A human study conducted in Hawaii was not considered adequate due to many
study design deficiencies (ATSDR, 1993c). 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,
1993c).
There are limited data on which to base an estimated exposure limit for
developmental effects. The NOEL in the two-generation study is not based on
sensitive endpoints and is only a factor of 3 removed from the LEL. The
developmental toxicity database is insufficient for heptachlor epoxide (per the IRIS
file). Consequently, the application of an uncertainty factor for the insufficiency of
the database may be necessary. The dog study, with a NOEL of 0.025 mg/kg/d,
can be used to calculate an exposure limit for developmental effects. The standard
uncertainty factors used in this calculation would typically take into consideration
inter- and intraspecies variability and a database factor.
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 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.
5.3.7.6 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, 1993). Six structurally related compounds have produced
tumors in mice and rats: chlordane, aldrin, dieldrin, heptachlor, and chlorendic acid
(IRIS, 1993).
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, 1993c).
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,
1993c).
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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
noted that preadolescent children may be more susceptible due to their greater
rate of glutathionine turnover (ATSDR, 1993c). In addition, children exposed
prenatally may be at higher risk, based on the results of developmental toxicity
studies.
See also a discussion of susceptibilities associated with pharmaceutical use in
Appendix C.
5.3.7.9 Interactive Effects—
See a discussion of organochlorine effects related to induction of the mixed
function oxidase system in Appendix C.
5.3.7.10 Critical Data Gaps—
The IRIS file notes data gaps as rat and rabbit teratology studies (IRIS, 1993). The
OPP notes the same data gaps (U.S. EPA, 1992c). 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, 1993C).
5.3,7.11 Summary of EPA Levels of Concern—
Chronic Toxicity
Carcinogenicity
5.3.7.12 Major Sources—
1.3x10'5mg/kg/d
9.1 per mg/kg/d.
ATSDR (1993c), IRIS (1993).
5.3.8 Hexachlorobenzene
5.3.8.1 Background—
Hexachlorobenzene is a byproduct of manufacturing and has been used as a
fungicide seed protectant in the past. It exists as a solid at ambient temperatures,
and in aquatic environments is found in higher quantities in sediment than water
due to its low solubility (ATSDR, 1990a).
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5.3.8.2 Pharmacokinetics—
Hexachlorobenzene is persistent in the body, accumulating preferentially in fat and
tissues with a high lipid content, due to its lipophilic nature. It is found in human
breast milk (ATSDR, 1990a), which may be a significant route of exposure for
young children.
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, 1990a). Based on animal studies,
the following systems are adversely affected following acute exposure: liver,
kidney, hematological, and dermal (ATSDR, 1990a). See also the discussion of
organochlorine pesticides in Appendix C.
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.
ATSDR has emphasized that the exposure estimates are unverified (ATSDR,
1990a).
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,
1990a). 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, based on animal studies, with a NOAEL of 0.08
mg/kg/d in a lifetime rat study. This has been converted by ATSDR to an MRL of
8 x 10'4 mg/kg/d using uncertainty factors of 10 each for inter- and intraspecies
variability (ATSDR, 1990a). This value is also the IRIS RfD for chronic systemic
toxicity (IRIS, 1993). Numerous other studies identified NOAELs in the same
numerical range. 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 (IRIS, 1993). It is not possible, based on the current data, to
determine whether the RfD will be protective against that effect.
The oral RfD of 8 x 10"4 mg/kg/d developed by IRIS and ATSDR was used to
calculate the fish consumption limits listed in Section 4 for chronic exposure
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toxicity. For a summary of the chronic systemic toxicity data, the reader is referred
to the Toxicity Profile for Hexachlorobenzene (ATSDR, 1990a).
5.3.8.5 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 study of nursing infants, the infants had blood levels of
hexachlorobenzene two to five times that of their mothers, as well as higher tissue
levels. A study of monkeys found that the concentration in milk was 17 times
higher than that in maternal serum (ATSDR, 1990a). 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, 1990a).
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, 1990a).
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, 1990a). 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, 1990a). Alterations in immune
function levels were reported in pre- and postnatally exposed rats at 4 mg/kq
(ATSDR, 1990a).
For purposes of quantitatively estimating an exposure limit, it is of concern that
prenatal and lactational exposure of humans at levels roughly estimated to be 0.7
to 2.9 mg/kg/d (maternal) induced serious structural changes in children and
increased mortality. Due to the poor quality of data supporting the exposure
estimates for the human exposure episode in Turkey and, more critically, the lack
of a no-effect level, it would be desirable to obtain a developmental study with a
more reliable exposure estimate. However, there do not appear to be such studies
currently available. Much higher exposure levels were required to cause structural
changes in experimental animals and the experimental results have not been
duplicated. These data suggest that humans may be more susceptible than the
animals studied.
In the absence of better data, the human study data from Turkey can be used to
calculate an estimated exposure limit for developmental effects. The standard
uncertainty factors used in this calculation would typically take into consideration
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intraspecies variability, the use of a LOAEL rather than a NOAEL, and the overall
inadequacy of the database. An additional modifying factor may be applied for the
poor quality of the exposure data and the severity of the effects noted at the
LOAEL. Due to the incomplete nature of the database and the other inadequacies
noted above, there would not be a high level of confidence in exposure limits
calculated from the current developmental toxicity database.
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, 1993).
Hexachlorobenzene was negative in dominant lethal studies (in vivo) at doses
from 60 to 221 mg/kg (ATSDR, 1990a).
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,
1993). The IRIS cancer potency of 1.6 per mg/kg/d was used to calculate the fish
consumption limits listed in Section 4 for carcinogenic effects.
Human studies have not yet yielded useful results. Followup 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, 1990a).
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, 1990a).
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Based on the toxicity data reviewed above, individuals with liver disease may be
at greater risk than the general population.
See also a discussion of susceptibilities associated with pharmaceutical use in
Appendix C.
5.3.8.9 Interactive Effects—
Hexachlorobenzene induces microsomal enzymes. See Appendix C for a
discussion of associated effects. 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 hexachloro-
benzene (ATSDR, 1990a).
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, 1990a). 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.
5.3.8.11 Summary of EPA Levels of Concern—
Chronic Toxicity
Carcinogenicity
5.3.8.12 Major Sources—
8x 10'4mg/kg/d
1.6 per mg/kg/d.
ATSDR (1990a), IRIS (1993).
5.3.9 Lindane (v-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.
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
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through urine as chlorophenols. The epoxide metabolite may be responsible for
carcinogenic and mutagenic effects (ATSDR, 1992b).
Male exposure to lindane through the environment results in accumulation in
testes and semen in addition to the tissues listed above (ATSDR, 1992b). See also
a discussion in Section 5.3.9.5 of the accumulation of lindane by pregnant women.
5.3.9.3 Acute Toxicity—
See the listing of usual effects associated with organochlorine exposure in
Appendix C. 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. ATSDR recommends an acute (14 days' exposure or less)
exposure MRL of 0.003 mg/kg/d based on neurotoxic effects in rats (ATSDR,
1992b).
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. Uncertainty factors of
10 each for inter- and intraspecies variability and the use of a less-than-lifetime
study were applied (IRIS, 1993). A recently completed 2-year study is under
evaluation and may provide additional information regarding toxicity (U.S. EPA,
1993k). Liver damage has been observed in animal studies (U.S. EPA, 1993k).
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, 1992b).
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 LEL of 2.5 mg/kg/d with 40 days of exposure in rats. No NOEL
was reported (ATSDR, 1992b).
Two recent reproductive studies in rats found adverse effects on the male
reproductive system. In a 7-week study, decreased sperm counts were noted at
50 mg/kg/d and, in a 180-day study, seminiferous tubular degeneration was noted
at 6 mg/kg/d with a NOEL of 3 mg/kg/d. An older study had identified the same
effects at 64.6 mg/kg/d in a 3-month 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-day study. Antiestrogenic effects were found at 20 mg/kg/d in female rats in a
15-week study with a NOEL of 5 mg/kg/d. This action was also found in two other
recent studies (ATSDR, 1992b). Based on current risk assessment methods, it
appears that the current IRIS RfD for chronic effects is protective against
reproductive system toxicity. However, the effects in both the male and female
reproductive systems have been evaluated only in short-term studies. Evaluation
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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.
It is not clear whether the IRIS RfD is protective against neurotoxic effects because
the study of behavioral effects that resulted in an LEL of 2.5 mg/kg/d was a short-
term study and no NOEL was identified. Neurotoxic effects have been reported
widely in human poisoning incidents and among occupationally exposed
individuals. Consequently, it is of significant interest for human toxicity. Additional
information is needed to determine whether the current RfD is protective against
neurotoxic effects.
5.3,9.5 Developmental Toxicity—
Two developmental toxicity studies in rats and rabbits both identified a NOEL 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, 1993k). 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, 1992b). 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 over which exposure occurred, with older women
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, 1992b). 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.
Multiple studies have reported that lindane exposure (as measured by body tissue
level of lindane) is associated with premature labor and spontaneous abortions.
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The causal relationship has not been established for this action (ATSDR, 1992b);
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.
Information was not located on developmental neurotoxicity, which may be an
expected effect of lindane based on the toxicity of other organochlorines. Based
on the limited data available, the most appropriate studies for use in calculating an
estimated exposure limit for developmental effects are the rat and mouse studies
that identified the development of extra ribs and fetal resorptions, respectively, at
an LEL of 5 mg/kg/d. Resorptions, which usually arise from early fetal death, are
typically the result of toxicity to the fetus. Information on the nature of that toxicity
was not available in the data reviewed for this document. In this case, the
resorptions could have arisen from systemic toxicity, or there may have been
hormonal effects that also jeopardized maintenance of pregnancy, as indicated by
the reproductive system toxicity data (see Section 5.3.9.4).
In estimating an exposure limit for lindane, the uncertainty generated by the
potential for lactation exposure must also be considered. It may be advisable to
use an additional modifying factor to account for lack of critical information in the
database regarding the actual dose at which toxic effects occurred (those more
sensitive than lethality), the potential for premature labor and spontaneous
abortions, and the potential for increased exposure via lactation. For purposes of
calculating an exposure limit, if the rat and mouse LEL were used, standard
uncertainty factors would typically take into consideration inter- and intraspecies
variability and the use of an LEL rather than a NOEL. A modifying factor may also
be applied. (See also Sections 5.3.9.8 and 5.3.9.9.)
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.
5.3.9.6 Mutagenicity—
In animals, ingestion of technical-grade hexachlorocyclohexane induced dominant
lethal mutations in mice. Studies found that iindane 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, 1992b).
5.3.9.7 Carcinogenicity—
Lindane has been classified as a probable/possible carcinogen (B2/C) based on
liver tumors in animals. The cancer potency is 1.3 per mg/kg/d (HEAST, 1992). In
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addition to tumors identified in experimental animals, human study data indicate
that this chemical may cause aplastic anemia (U.S. EPA, 1993a). Lindane is
currently under review by EPA. Lindane's related isomers, alpha and beta
hexachlorocyclohexane, are also classified as probable human carcinogens and
have cancer potencies similar to that of lindane. The cancer potency obtained from
the HEAST tables was used to calculate the fish consumption limits listed in
Section 4 for carcinogenic effects.
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, 1992b). People with epilepsy, cerebrovascular accidents, or
head injuries who have lower thresholds for convulsions may be at greater risk of
lindane-induced CNS 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, 1992b).
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 immunosup-
pression because their immune systems do not reach maturity until
10 to 12 years of age (ATSDR, 1990b).
See also a discussion of susceptibilities associated with pharmaceutical use in
Appendix C.
5.3.9.9 Interactive Effects—
See a discussion of organochlorine effects related to induction of the mixed
function oxidase system in Appendix C.
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, 3, and 5
hexachlorocyclohexane isomers may reduce the neurotoxic effects of lindane
(ATSDR, 1992b).
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 for these combinations
in rodents exposed via gavage (MIXTOX, 1992).
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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. A large group of
international studies recently submitted to ATSDR are currently under review and
six studies are ongoing in the United States (ATSDR, 1992b).
5.3.9.11 Summary of EPA Levels of Concern—
Chronic Toxicity
Carcinogenicity
3x 10"4mg/kg/d
1.3 per mg/kg/d.
5.3.9.12 Major Sources—
ATSDR (1992b), HSDB (1993), IRIS (1993).
5.3.10 Mirex
5.3.10.1 Background—
Mirex is a polymerizing agent and was used as an organochlorine pesticide and
fire retardant until 1975 (U.S. EPA, 1993a). Mirex has the potential to concentrate
many thousand-fold 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, 1993o).
No clear data on half-life in humans was 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.
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5.3.10.3 Acute Toxicity—
See the listing of usual effects associated with organochlorine exposure in
Appendix C. 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 endo-
plasmic reticulum (U.S. EPA, 1993o).
5.3.10.4 Chronic Toxicity—
IRIS lists a chronic exposure RfD of 2 x 10"4mg/kg/d for mirex based on a NOAEL
of 0.07 mg/kg/d from a chronic dietary rat study. The IRIS file notes that the
previous RfD was 2 x 10'6 mg/kg/d. The IRIS file states that a dose-related
increase in hyperplasia of the parathyroid gland was observed in males in the
critical study at and above 0.007 mg/kg/d (IRIS, 1993). It is not stated in the file
why this value was not used as a LOAEL; although it is noted that the effect was
not observed in other studies. Additional effects noted in the study were:
nephropathy, renal medullary hyperplasia, multiple types of liver damage, splenic
fibrosis, and cystic follicles of the thyroid. The RfD is based on the latter two
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 testis (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, 1993).
A subchronic study in rats identified an LEL of 0.01 mg/kg/d with liver lesions and
thyroid injury at the lowest dose tested. Another subchronic study in rats noted
similar effects with a NOEL of 0.01 in which only females were tested. A 21-month
rat feeding study identified an LEL of 0.01 mg/kg/d based on histological lesions
in the liver and thyroid and altered enzyme levels (U.S. EPA, 1993o). These
results are supported by 28-day feeding studies, which identified LELs at a similar
level to the studies listed above. Histopathology of the liver was noted at 0.025
mg/kg/d in two rat studies. No NOELs were identified (U.S. EPA, 1993o). Both
structural and functional adverse effects on the thyroid have been observed in
experimental animals. The effects persisted for more than 1 year after treatment
ceased. Neurobehavioral effects have also been associated with mirex exposure
(Hayes and Laws, 1991).
Both the longer-term and the subchronic studies, which identified LELs of 0.01
mg/kg/d, suggest that toxicity occurs at levels below those identified in the NTP
study, which is used as the basis for the IRIS RfD. The lower LELs can be used
to calculate an alternative estimated exposure limit. The standard uncertainty
factors used in this calculation would typically take into consideration inter- and
intraspecies variability and the use of an LEL rather than a NOAEL.
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5.3.10.5 Developmental Toxicity—
Numerous developmental toxicity studies have been conducted on mirex. Effects
associated with exposure include undescended testes (U.S. EPA, 1993o), fetal
cataracts, edema, ectopic gonads, hydrocephaly, abnormal kidney enzyme levels,
decreased brain and liver weights, scoliosis, runts, cleft palates, heart defects
(anatomical), effects on the fetal electrocardiogram, and decreased fertility.
Multiple studies have noted increased mortality and cataracts resulting from both
pre- and postnatal exposure (IRIS, 1993). Many of the deaths have been due to
congestive heart failure. Studies utilizing cross-fostering of litters have found that
both cataracts and reduced viability were less pronounced when neonates did not
receive their mothers' milk (mothers were dosed with mirex during pregnancy)
(Hayes and Laws, 1991).
Mirex is lipophilic and has been found in humap breast milk. A small study in the
Great Lakes region found levels from 0.1 to 0.6 ppm in human milk (Hayes and
Laws, 1991).
Many of the developmental studies noted significant effects at the lowest dose
tested (e.g., cardiac function abnormalities occurred at the LOEL of 0.25 mg/kg/d)
(IRIS, 1993). Consequently, they do not provide a useful threshold for effects. One
study of cardiac effects in the fetus found a dose-related increase in first-degree
heart block; however, the levels tested were quite high (5, 6, 7, and 10 mg/kg). A
high rate of stillbirth and postnatal mortality, first- and second-degree heart blocks,
respiratory distress, and cataracts was observed in a prenatal exposure study with
an LEL of 1 mg/kg/d (U.S. EPA, 1993o).
A single dose of 1.25 mg/kg to pregnant rats caused reduced viability and growth
and a high incidence of cataracts in offspring. A developmental study in rats
identified an LEL of 0.25 mg/kg/d with an increase in stillborn pups and decreased
viability. A one-generation mouse study identified an LEL of 0.075 mg/kg in a
single dose causing reduced litter size; no NOEL was identified. A one-generation
rat study identified an LEL of 0.125 mg/kg/d with decreased litter size, histopatho-
logical changes in the liver and thyroid, and cataracts (U.S. EPA, 1993o). Bio-
chemical alterations include significant decreases in plasma protein concentrations
and colloid osmotic pressure in fetuses (U.S. EPA, 1993o).
Mirex causes serious adverse effects in multiple organ systems in developing
animals. Frank teratogenic effects are observed at levels that are much higher
than those required to produce other effects. Teratogenic effects have been
observed at an LEL of 6.0 mg/kg with an increased incidence of visceral anomalies
and deciduomas. A teratogenic NOEL of 3.0 mg/kg was identified (U.S. EPA,
1993o).
Due to the absence of a NOEL for many of the effects, there is considerable
uncertainty regarding the calculation of an exposure limit for developmental
effects. However, the seriousness of the effects argues for inclusion of a devel-
opmental risk value to provide some guidance for exposure. Based on the
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information reviewed, the most sensitive species appears to be the mouse, with
effects observed at 0.075 mg/kg in a single dose (LEL). Studies in other species
identified multiple effects at exposure levels equal to, or less than, a factor of
twofold greater (0.125 and 0.25 mg/kg/d). If the mouse study results are 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 LEL rather than a NOAEL. Due to the multiple and
serious effects associated with mirex exposure during development, and the
potential for bioaccumulation and exposure through the placenta and via breast
milk, an additional modifying factor may be applied.
As noted above, mirex 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 mirex, even if exposure is
reduced during pregnancy, the outcome of that pregnancy may be affected,
depending on the timing and extent of prior exposure.
5.3.10.6 Mutagenicity—
Most genotoxicity tests reported in the tox one-liners are bacterial assays and are
negative (U.S. EPA, 1993o). 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. Additional information is needed
on the nature of the toxicity.
5.3.10.7 Carcinogenicity—
Mirex has been classified as a probable human carcinogen (B2) based on liver
and adrenal tumors in experimental animals. The cancer potency is 1.8 per
mg/kg/d (HEAST, 1995). This chemical is currently under review by EPA.
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.
See also a discussion of susceptibilities associated with pharmaceutical use in
Appendix C.
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5.3.10.9 Interactive Effects—
See a discussion of organochlorine effects related to induction of the mixed
function oxidase system in Appendix C. No additional data were located.
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 NOEL 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 NOELs for sensitive endpoints in this area is needed.
5.3.10.11 Summary of EPA Levels of Concern-
Chronic Toxicity
Carcinogenicity
2x10"4mg/kg/d
1.8 per mg/kg/d.
5.3.10.12 Major Sources—
Hayes and Laws (1991), IRIS (1993), U.S. EPA (1993o).
5.3.11 Toxaphene
5.3.11.1 Background—
Toxaphene is an organochlorine pesticide that is comprised of a mixture of 670
chlorinated camphenes. It was banned for most uses in 1982; 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—
Toxaphene is rapidly degraded 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 (ATSDR, 1990b).
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Some adverse effects of toxaphene may result from repeated exposure that do not
occur with a single exposure to a lesser dose. 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. This is significant when considering potential risks arising from
chronic exposure to toxaphene (ATSDR, 1990b).
5.3.11.3 Acute Toxicity—
Acute high-level exposures to toxaphene have resulted in death in adults and
children with an estimated minimum lethaLdose of 2 to 7 g, which is equivalent to
29 to 100 mg/kg for an adult male. Long-term damage to the central nervous
system and liver has also been observed. The kidney and adrenal glands are also
target organs (ATSDR, 1990b). A 1 -day NOAEL of 10 mg/kg/d is available from a
dog study that used death as the effect of concern. A 14-day LOAEL of 5 mg/kg/d
was identified in an 8-day study that was used as the basis for an MRL for acute
exposure of 0.005 mg/kg/d by ATSDR (ATSDR, 1990b). See the listing of usual
effects associated with organochlorine exposure in Appendix C.
5.3.11.4 Chronic Toxicity—
IRIS does not provide a discussion of chronic effects of exposure to toxaphene or
an RfD (IRIS, 1993). The EPA Office of Water and Office of Pesticide Programs
have calculated an RfD of 3.6 x 10'4 mg/kg/d based on the absence of liver,
kidney, and thyroid effects in rats exposed to toxaphene via the oral route for 26
weeks.
Chronic exposure to toxaphene may result in damage to the following systems:
liver, kidney, adrenal, immunological, and neurological. The use of the liver as the
endpoint of concern is supported by a recent subchronic oral rat study that found
NOAELs of 0.28 for males and 0.38 for females with liver and kidney effects
(ATSDR, 1990b).
Chronic exposure to toxaphene may cause hormonal alterations. A study found
increased levels of hepatic metabolism in vivo and in vitro of estradiol and estrone
and a decrease in their uterotropic action. Duration of exposure was not specified
in the source reviewed (HSDB, 1993). See also notes regarding estrogenic activity
in Section 5.3.11.7.
5.3.11.5 Developmental Toxicity—
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 with a LOAEL of 1.5 mg/kg/d. Impairment
of behavioral maturation (e.g., reflexes) occurred at 0.05 mg/kg/d in a rat study
with 47 days of exposure. Delayed ossification (bone development) and alterations
in kidney and liver enzymes suggestive of organ-specific toxicity were observed
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at 15 mg/kg/d. Other adverse effects noted in offspring of maternally exposed
individuals included histological changes in the liver, thyroid, and kidney (ATSDR,
1990b).
Women exposed to toxaphene by entering a field that had recently been sprayed
exhibited a higher incidence of chromosomal aberrations in cultured lymphocytes
than was found in unexposed women. Dermal and inhalation were the probable
routes of exposure; however, the exposure was not quantified (ATSDR, 1990b).
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, 1990b).
However, chromosomal aberrations (as observed in women) would lead to
decreased fertility due to early fetal loss and may result in heritable birth defects.
Toxaphene is known to be conveyed into milk rapidly after maternal exposure to
the chemical. The half-life of toxaphene has been estimated at 9 days. It has been
found in the milk of cows at all doses tested (20 to 40 ppm). In cows exposed to
20 to 140 ppm in food (mg/kg/d conversion not available) for 8 weeks, milk
concentrations increased rapidly; they decreased rapidly following cessation of
exposure. Information was provided on the relationship between feed and milk
concentrations. The exposure range was from 20 ppm feed with milk
concentrations reaching 0.36 ppm to 140 ppm feed with a maximum of 1.89 ppm
in milk (ATSDR, 1990b). It may be advisable to use these data to estimate the
human dose to nursing infants.
Other aspects of developmental toxicity associated with toxaphene are based on
effects observed in adult individuals that are known to pose higher risks to
children. The ATSDR has cautioned that:
embryos, fetuses, and neonates up to age 2 to 3 months may be at
increased risk of adverse effects . . . because their enzyme
detoxification systems are immature
and
Infants and children are especially susceptible to immunosup-
pression because their immune systems do not reach maturity until
10 to 12 years of age (ATSDR, 1990b).
Immunosuppression was noted in multiple subchronic exposure animal studies.
ATSDR also noted that:
animal studies suggest that detoxification of the toxaphene mixture
may be less efficient in the immature human than the metabolism
and detoxification of the single components such as Toxicant A or
B (ATSDR, 1990b).
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5.3 ORGANOCHLORINE PESTICIDES
ATSDR provides an MRL for intermediate exposures (14 to 365 days) of 5 x 10~5
mg/kg/d based on an LEL of 0.05 mg/kg/d associated with impaired behavioral
development. Uncertainty factors of 10 each for inter- and intraspecies variability
and the use of an LEL were applied (ATSDR, 1990b). This value can be used as
the estimated exposure limit for developmental effects in developing fish
consumption limits.
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. 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 toxaphene, even if exposure
is reduced during pregnancy, the outcome of that pregnancy may be affected,
depending on the timing and extent of prior exposure.
5.3.11.6 Mutagenicity-—
There are 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 is not known and 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 mutagenic than the nonpolar fraction
(ATSDR, 1990b; HSDB, 1993).
Changes in human genetic material have been noted in workers exposed to
toxaphene (HSDB, 1993).
5.3.11.7 Carcinogenicity—
Toxaphene is classified as a probable human carcinogen (B2) by EPA based on
oral studies in animals. The cancer potency is 1.1 per mg/kg/d, based on liver
tumors in experimental animals (IRIS, 1992). This value was used to calculate fish
consumption limits listed in Section 4 for carcinogenic effects. No conclusive
human epidemiological studies are available for toxaphene (ATSDR, 1990b).
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 et al., 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, 1990b). Because this
information was obtained from an LD50 study, it cannot be used directly to modify
risk values. The Centers for Disease Control has specified that:
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This has important implications with regard to the possible
increased susceptibility of humans who ingest a protein-deficient
diet and live in areas of potential exposure to toxaphene. (ATSDR,
1990b)
If a population with protein-deficient diets is the target group for fish consumption
limits, an additional modifying factor of 10 could be used in determining the
appropriate exposure limit for developmental effects. A factor of 10 rather than 3
(observed in animal studies) is recommended because it is unknown whether
human populations exposed under the same conditions will be more or less
susceptible than the animals tested and because the results were obtained from
an LD50 study rather than a study with a more sensitive toxic endpoint.
The nervous system is a primary target of toxaphene toxicity. Individuals with
latent or clinical neurological diseases, such as epilepsy or behavioral disorders,
may be at higher risk. In addition, children may be especially susceptible to
toxaphene-induced neurotoxicity based on early reports of acute ingestion toxicity
(ATSDR, 1990b).
As discussed in Section 5.3.11.5, ATSDR has identified pregnant women, fetuses,
infants, and children as populations at greater risk. 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).
See also a discussion of susceptibilities associated with pharmaceutical use in
Appendix C.
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 due to its increased metabolism arising from toxaphene's
induction of microsomal enzymes (ATSDR, 1990b).
See a discussion of organochiorine effects related to induction of the mixed
function oxidase system in Appendix C.
Based on acute studies 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, reduces the acute toxicity of
chemicals that require this system for detoxification (ATSDR, 1990b).
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In experimental animals, toxaphene antagonized the tumorigenic activity of
benzo(a)pyrene in the lung. It was theorized that this occurred because toxaphene
inhibited the biotransformation of B(a)P to a reactive metabolite or by promoting
its degradation to nonactive forms (ATSDR, 1990b).
MIXTOX has reported synergism between chlordane, toxaphene, and malathion
in mice exposed via gavage and additive interactions between chlordane and
toxaphene. Antagonism was reported between toxaphene and diazinon in rats
exposed via gavage. Mixed results have been obtained between lindane and
toxaphene (MIXTOX, 1992).
5.3.11.10 Critical Data Gaps—
The following data gaps have been identified by ATSDR, EPA, and NTP:
mammalian germ cell genotoxicity, studies that investigate sensitive develop-
mental 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, 1990b).
5.3.11.11 Summary of EPA Levels of Concern-
Chronic Toxicity
Carcinogenicity
5.3.11.12 Major Sources—
3.6x1(r4mg/kg/d
1.1 permg/kg/d.
ATSDR (1990b), HSDB (1993), IRIS (1993).
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5.4 ORGANOPHOSPHATE PESTICIDES
5.4 ORGANOPHOSPHATE PESTICIDES
In addition to the discussions of individual target analytes, please refer to the
discussion of toxicity characteristics of the organophosphate chemical group in
Appendix C.
5.4.1 Chlorpyrifos
5.4.1.1 Background—
Chlorpyrifos is an organophosphate insecticide that is applied throughout the
United States for various agricultural uses.
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,
1992g). Chlorpyrifos is rapidly metabolized and excreted based on studies in
animals (Hayes and Laws, 1991).
5.4:1.3 Acute Toxicity-
See the listing of usual effects associated with organophosphate exposure in
Appendix C.
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, 1993). 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 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. Problems related to the use of cholines-
terase inhibition as a critical endpoint are discussed in Appendix C. The value
listed on IRIS was confirmed in 1993 by an Office of Pesticide Programs RfD Peer-
Review Committee (U.S. EPA, 1993e).
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
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5.4 ORGANOPHOSPHATE PESTICIDES
subchronic study at higher doses, the same effects were observed along with
increased brain and heart weight (U.S. EPA, 1992g).
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 Developmental Toxicity—
Chlorpyrifos is fetotoxic in numerous species. In a 1987 rat study, a developmental
toxicity NOEL of 2.5 mg/kg/d was determined; at the LOEL of 15 mg/kg/d
postimplantation losses were observed. In a 1991 rat study, which was rated as
"guideline" by OPP, a developmental NOEL of 1 mg/kg/d was obtained, with
increased pup mortality at 5 mg/kg/d. No data were available for this document on
underlying causes of mortality. Decreased fetal length and increased skeletal
variants were noted in mice with a fetotoxic NOEL for the study of 10 mg/kg/d. In
a 1987 study on rabbits, increased skeletal variants and an increased incidence
of unossified sternebra and xiphisternum were observed at 81 mq/kq/d (U S EPA
1992g).
A1991 rat study yielded the most conservative NOEL at 1 mg/kg/d. Unfortunately,
the observed endpoint was mortality. An evaluation of the underlying causes of
mortality may yield a more sensitive endpoint. If this study were used to estimate
an exposure limit for developmental effects, the standard uncertainty factors used
in this calculation would typically take into consideration inter- and intraspecies
variability and data gaps, based on the inability of the current studies to identify
critical information.
Currently available data regarding developmental toxicity are limited because
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. In cases such as this, where the available
studies provide information only on gross measures of toxicity (i.e., death), it may
be advisable to use the RfD for chronic toxicity and consider modifications for
application to pregnant women and children.
5.4.1.6 Mutagenicity—
The results of mutagenicity assays of chlorpyrifos are mixed. Chlorpyrifos was
weakly positive with and without activation in gene conversion and recombination
assays and positive for direct damage to DNA in B. subtilis (U.S. EPA, 1992g). In
vivo assays of mouse liver DNA and RNA indicated that chlorpyrifos caused more
DNA and RNA alkyiation than other organophosphates (HSDB, 1993). Its toxicity
is probably related to formation of its oxon analog (chlorpyrifosoxon) and
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5.4 ORGANOPHOSPHATE PESTICIDES
subsequent enzyme inhibition of cholinesterase activity, carboxylases, and
mitochondrial oxidative phosphorylases.
5.4.1.7 Carcinogenicity-
Insufficient information is available to determine the carcinogenic status of
chlorpyrifos.
5.4.1.8 Special Susceptibilities-
See a discussion of susceptibilities associated with organophosphate exposure in
Appendix C.
5.4.1.9 Interactive Effects-
No data were located.
5.4.1.10 Critical Data Gaps—
IRIS lists the following data gap: chronic feeding/oncogenicity study in rats (IRIS,
1993). Additional data are needed on the noncholinesterase effects of chronic
exposure and on the toxicity that underlies early pup mortality in developmental
studies.
5.4.1.11 Summary of EPA Levels of Concern-
Chronic Toxicity 3 x 10"3 mg/kg/d
Carcinogenicity Insufficient data to determine carcinogenic status.
5.4.1.12 Major Sources—
HSDB (1993), IRIS (1993), U.S. EPA (1992g).
5.4.2 Diazinon
5.4.2.1 Background—
Diazinon is an organophosphorus insecticide that has been widely used since its
introduction in 1952.
5.4.2.2 Pharmacokinetics—
Very little data were located. Metabolism appears to proceed by similar but
somewhat different paths in various mammalian species (HSDB, 1993). Human
milk may contain trace amounts of diazinon based on the results of exposure in
cows (HSDB, 1993).
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5.4 ORGANOPHOSPHATE PESTICIDES
5.4.2.3 Acute Toxicity—
Diazinon is highly toxic. The estimated adult oral fatal dose is approximately 25 g
(HSDB, 1993). See the listing of usual effects associated with organophosphate
exposure in Appendix C.
5.4.2.4 Chronic Toxicity—
IRIS does not currently provide an oral RfD because it is under review within the
Agency (IRIS, 1993). OPP provides an RfD of 9 x 10'5 mg/kg/d based upon
cholinesterase inhibition observed in a 90-day rat feeding study with a NOEL of
0.009 mg/kg/d and uncertainty factors totaling 100 (U.S. EPA, 1992d). Problems
related to the use of cholinesterase inhibition as a critical endpoint are discussed
in Appendix C.
Very little dose-response data are available on chronic systemic toxicity, other than
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,
1993f). 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 Developmental Toxicity—
The reproductive/teratogenic studies listed in the tox one-liners report no adverse
effects at the highest doses tested (U.S. EPA, 1993f).
HSDB reported multiple studies indicating diazinon is teratogenic. 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 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 (Hayes,
1982). This study appears to provide a relatively sensitive endpoint for evaluation
of developmental effects associated with exposure to diazinon. If the LOAEL of
0.18 mg/kg/d were used, the uncertainty factors would typically take into
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5.4 ORGANOPHOSPHATE PESTICIDES
consideration inter- and intraspecies variability and the use of a LOAEL rather than
a NOAEL Based on the available information, the current IRIS RfD would be
protective against developmental toxicity.
5.4.2.6 Mutagenicity—
Most mutagenicity assays were negative; one positive sister chromatid exchange
assay was noted (U.S. EPA, 1993f). A study on the effect of diazinon on mitosis
in human lymphocytes reported chromosomal aberrations in 74 percent of the cells
at0.5mg/mL(HSDB, 1993).
5.4.2.7 Carcinogenicity—
Insufficient information is available to determine the carcinogenic status of
diazinon.
5.4.2.8 Special Susceptibilities
See a discussion of susceptibilities associated with organophosphate exposure in
Appendix C.
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
doses and defined a NOAEL would be useful in establishing an appropriate RfD.
5.4.2.11 Summary of EPA Levels of Concern-
Chronic Toxicity
Carcinogenicity
9 x 10'5 mg/kg/d based on cholinesterase inhibition
Insufficient information to determine carcinogenic status.
5.4.2.12 Major Sources-
Hayes (1982), HSDB (1993), U.S. EPA (1993f).
5.4.3 Disulfoton (disyston)
5.4.3.1 Background—
Disulfoton is an organophosphate pesticide with high acute toxicity to all mammals.
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5.4 ORGANOPHOSPHATE PESTJCJDES
5.4.3.2 Pharmacokinetics—
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, 1993h). 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 (HSDB
1993).
5.4.3.3 Acute Toxicity-
See the listing of usual effects associated with organophosphate exposure in
Appendix C. The acute oral LD50 in animals ranges from 2 to 27.5 mg/kg (U.S.
EPA, 1993h). 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 LEL of 0.04 mg/kg/d from
a 2-year rat study that was associated with cholinesterase inhibition and optic
nerve degeneration (IRIS, 1993). The IRIS RfD was calculated using a modifying
factor of 10 to account for possible findings in the additional recommended optic
toxicity studies (U.S. EPA, 1992c). Standard methods would typically utilize 10
each for inter- and intraspecies variability and for the use of an LEL rather than a
NOEL. This plus the modifying factor of 10 would yield an RfD of 4 x 10'6 mg/kg/d.
Although there is some question regarding the use of cholinesterase inhibition as
the basis for establishing an RfD (problems related to the use of cholinesterase
inhibition as a critical endpoint are discussed in Appendix C), optic effects also
serve as the basis for this RfD (U.S. EPA, 1992d).
Numerous other effects of disulfoton have been reported at doses within 1 order
of magnitude of the LEL identified in the critical study. Significant toxicity in multiple
organ systems has been observed at 0.1 mg/kg/d (the lowest dose tested) for the
following systems: spleen, liver, pituitary, brain, seminal vesicles, and kidneys
(IRIS, 1993). 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). Based on the chronic exposure information reviewed and
standard assumptions regarding the use of uncertainty factors, the IRIS RfD
appears to be protective against the effects listed above.
5.4.3.5 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
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5.4 ORGANOPHOSPHATE PESTICIDES
stages of liver damage at 0.5 mg/kg/d. No NOEL 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, 1993, and U.S. EPA, 1993h).
Toxicity was incompletely characterized in this study, and additional studies are
needed to provide an adequately defined NOEL for developmental effects.
Because multiple serious effects were observed at the lowest dose tested, the
multigeneration study does not provide an optimal basis for calculation of an
exposure limit for developmental effects. However, it does indicate that adverse
developmental effects may occur with exposure to disulfoton and provides greater
detail on these effects than do the other studies available.
A more recent two-generation rat study identified a NOEL of 0.04 mg/kg/d with an
LEL of 0.12 mg/kg/d based on decreased litter sizes, pup survival, and pup
weights at the LEL (U.S. EPA, 1993h). This study does not appear to provide the
same level of analysis of sensitive endpoints as the three-generation study
discussed above. However, it identifies a lower NOEL and LEL than the two older
studies. If this study is used to calculate an estimated exposure limit for
developmental effects, the uncertainty factors typically used in this calculation
would take into consideration inter- and intraspecies variability. A modifying factor
could be used for the lack of data on the level at which toxicity occurred that led
to death. Additional studies are needed to identify the NOEL for sensitive
measures of the testicular, liver, and kidney toxicity identified in the multigeneration
study.
5.4.3.6 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, 1993h), and in other in vitro assays
(HSDB, 1993).
5.4.3.7 Carcinogenicity—
Insufficient information is available to determine the carcinogenic status of
disulfoton.
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. Also
see a discussion of susceptibilities associated with organophosphate exposure in
Appendix C.
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5.4 ORGANOPHOSPHATE PESTfCfDES
5.4.3.9 Interactive Effects-
No data were located.
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, 1993). HSDB notes that, because
of data gaps, a full risk assessment cannot be completed. Major relevant data
gaps noted under the 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 Levels of Concern-
Chronic Toxicity
Carcinogenicity
5.4.3.12 Major Sources—
4 x1Q-5 mg/kg/d
Insufficient data to determine carcinogenic status.
HSDB (1993), IRIS (1993), U.S. EPA (1993h).
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—
No data were located.
5.4.4.3 Acute Toxicity—
See the listing of usual effects associated with organophosphate exposure in
Appendix C.
5.4.4.4 Chronic Toxicity—
A 1970 study of 10 men (six test subjects) with a NOEL of 0.05 mg/kg/d found
plasmerand brain cholinesterase inhibition (IRIS, 1993). IRIS provides an RfD of
5x10 mg/kg/d based on a subchronic study in dogs that found a NOEL of 0.06
and 0.07 mg/kg/d for males and females, respectively, with the same effects as the
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5.4 ORGANOPHOSPHATE PESTICIDES
human study. Uncertainty factors of 10 each for intraspecies sensitivity and for the
effects observed at 0.71 mg/kg/d in the dog study (IRIS, 1993). Problems related
to the use of cholinesterase inhibition as a critical endpoint are discussed in
Appendix C.
5.4.4.5 Developmental Toxicity—
A developmental NOEL of 0.6 mg/kg/d was obtained in a rat study that found
delayed ossification at an LEL of 2.4 mg/kg/d (IRIS, 1993). A rabbit study by the
same laboratory also identified an LEL of 2.4 mg/kg/d with an increased incidence
of fused sternal centers and fetal resorptions at that dose level. The NOEL was 0.6
mg/kg/d (U.S. EPA, 1993n). A three-generation rat study was also listed in the tox
one-liners; however, information was provided only on cholinesterase inhibition
levels (U.S. EPA, 1993n).
The NOEL of 0.6 mg/kg/d from the rat and rabbit studies can be used to calculate
an estimated exposure limit for developmental effects. The uncertainty factors
would typically take into consideration inter- and intraspecies variability.
Teratogenic effects and fetal death often occur at exposure levels considerably
higher than levels associated with systemic toxicity. Other.organophosphates have
shown this gradient of effects (see diazinon, Section 5.4.2). Consequently, there
is concern that the studies available for evaluation may not fully characterize the
developmental toxicity of ethion.
In cases such as this, where the available studies provide information on only
gross measures of toxicity (i.e., death), it may be advisable to use the RfD for
chronic toxicity and consider modifications for application to pregnant women and
children.
5.4.4.6 Mutagenlcity—
The tox one-liners listed no positive study results.
5.4.4.7 Carcinogenicity—
Insufficient information is available to determine the carcinogenic status of ethion.
5.4.4.8 Special Susceptibilities-
See a discussion of susceptibilities associated with organophosphate exposure in
Appendix C.
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, 1993n).
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5.4 ORGANOPHOSPHATE PESTICIDES
5.4.4.10 Critical Data Gaps—
IRIS lists a chronic dog feeding study as a data gap (IRIS, 1993). A multigenera-
tion study and a developmental study that evaluate neurobehavioral toxicity are
needed to clarify developmental effects.
5.4.4.11 Summary of EPA Risk Values—
Chronic Toxicity 5x 10'4 mg/kg/d
Carcinogenicity Insufficient data to determine carcinogenic status.
5.4.4.12 Major Sources—
IRIS (1993), U.S. EPA (1993n).
5.4.5 Terbufos
5.4.5.1 Background—
Terbufos is an organophosphorus insecticide.
5.4.5.2 Pharmacokinetics—
No data were located.
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 1992f). See the listing of usual effects
associated with organophosphate exposure in Appendix C.
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. IRIS does not provide an
RfD for terbufos. HEAST lists an RfD of 2.5 x 1CT5 mg/kg/d based on
cholinesterase inhibition in a 6-month dietary dog study with a NOEL of 0.0025
mg/kg/d. Uncertainty factors of 10 each for inter- and intraspecies variation were
used. No uncertainty factor was used for the subchronic nature of the study. The
HEAST table states that this value is under review (HEAST, 1992). OPP has
calculated an RfD of 1.3 x 10'4 mg/kg/d for terbufos (U.S. EPA, 1996b).
Quantitative chronic toxicity information on cholinesterase inhibition is available.
In rats, a 1974 lifetime oral study found a LOEL of 0.0125 mg/kg/d (the lowest
dose tested); a 1987 1-year oral study found a NOEL of 0.025 mg/kg/d. In dogs,
a 1972 6-month oral study found a NOEL of 0.0025 mg/kg; a 1986 1-year study
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5.4 ORGANOPHOSPHATE PESTICIDES
found a LOEL of 0.015 mg/kg/d (the lowest dose tested); a 1987 28-day dog study
identified a NOEL of 0.00125 mg/kg/d (U.S. EPA, 1992f).
Quantitative data on chronic effects that are not directly related to cholinesterase
inhibition are limited, due to 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, 1992f). 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-month) rat study (animals were not examined for this lesion at lower
exposure levels) (U.S. EPA, 1992f).
5.4.5.5 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 NOEL of 0.1
mg/kg/d with increased fetal resorptions at 0.2 mg/kg/d; a 1988 rabbit study
identified a NOEL of 0.25 mg/kg/d with fetal resorptions at 0.5 mg/kg/d. A 1973
multigeneration reproductive study found a NOEL of 0.0125 mg/kg/d in rats, based
on an increase in the percentage of deaths in offspring (U.S. EPA, 1992f).
The increase in deaths in offspring in the multigeneration study does not provide
insight into the causes of death. Fetal resorptions, noted in the developmental
studies, often result from gross abnormalities leading to early fetal death or from
direct fetotoxicity. A NOEL for adverse effects would be a preferable endpoint, with
exploration of the causes underlying fetal loss. A multigeneration study that
evaluated sensitive endpoints, including ocular effects and liver and lymph node
toxicity, which were observed at low doses in adults animals (see Section 5.4.5.4),
would provide a better basis for determining a safe developmental exposure level.
Based on the developmental data currently available, the multigeneration study
NOEL of 0.0125 appears to be the most sensitive study (perhaps because
exposure occurred over a longer period of time than in the other developmental
toxicity studies). However, the developmental toxicity database embodies con-
siderable uncertainty. Many chronic exposure effects were observed at levels
approximating the NOEL obtained from the multigeneration study. Although these
effects were not reported in the developmental toxicity studies, it is not known
whether effects observed in adult studies, such as liver and lymph node toxicity,
were evaluated in the developmental studies. Postnatal exposure would be
expected to be at least as toxic to young individuals as to adults.
If the multigeneration study is used to calculate an exposure limit for devel-
opmental effects, the standard uncertainty factors would typically take into
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5.4 ORGANOPHOSPHATE PESTICIDES
consideration intra- and interspecies variability and the inadequacy of the data-
base. In cases such as this, where the available studies provide information on
only gross measures of toxicity (i.e., death), it may be advisable to use the RfD for
chronic toxicity and consider modifications for application to pregnant women and
children.
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, 1992f). '.
5.4.5.7 Carcinogenicity—
Insufficient information is available to determine the carcinogenic status of
terbufos.
All oncogenicity tests on terbufos have been considered negative by OPP (U.S.
EPA, 1992f). However, further exploration of mesenteric and mandibular lymph
node hyperplasia identified in a 3-month study (noted above) is warranted because
hyperplasia is often a precancerous condition. Evaluation of this endpoint in a
lifetime study is necessary to determine the ultimate course of the hyperplasia.
5.4.5.8 Special Susceptibilities-
See a discussion of susceptibilities associated with organophosphate exposure in
Appendix C.
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, 1992f).
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 NOEL 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
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5.4 ORGANOPHOSPHATE PESTICIDES
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 Levels of Concern-
Chronic Toxicity 1.3 x 10'4 mg/kg/d
Carcinogenicity Insufficient data to determine carcinogenic status.
5.4.5.12 Major Sources—
HSDB (1993), U.S. EPA (1992f).
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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 RfD of 3 x 10'3 mg/kg/d based on a NOAEL of 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, 1993).
5.5.1.5 Developmental Toxicity—
A three-generation rat study provided a NOEL of 0.5 mg/kg/d and an LEL of 5
mg/kg/d. A rat teratology study identified a fetotoxic NOEL 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,
1993, U.S. EPA, 19931). A rabbit teratology study data gap is noted in the IRIS file
(IRIS, 1993). Nitrofen, a close structural relative of oxyfluorfen has been studied
more extensively. Studies of nitrofen identified multiple varied developmental
abnormalities associated with prenatal exposure (Hayes and Laws, 1991).
The multigeneration study is the most sensitive study of those reviewed; this may
be due to the longer period of exposure and followup than the prenatal exposure
studies. The standard uncertainty factors used in this calculation would typically
take into consideration inter- and intraspecies variability. Additional information is
needed on the nature of effects at the LEL.
5.5.1.6 Mutagenicity—
Results of mutagenicity assays on oxyfluorfen are mixed (U.S. EPA, 19931).
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5.5 CHLOROPHENOXY HERBICIDES
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 potency of 0.13 is
provided by OPP (U.S. EPA, 1992d).
5.5.1.8 Interactive Effects-
No data were located.
5.5.1.9 Critical Data Gaps—
The IRIS file notes a rabbit teratology study as a data gap.
5.5.1.10 Summary of EPA Levels of Concern-
Chronic Toxicity 3 x 10"3 mg/kg/d
Carcinogenicity 0.13 per mg/kg/d.
5.5.1.11 Major Sources—
IRIS (1993), U.S. EPA (1993I).
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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 (U.S.
EPA, 1995). 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, 1995). ATSDR (1995) 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
Benzofjbjfluoranthene
Benzo[/f]fluoranthene
Benzo[/]fluoranthene
Benzo[gr,/7,/]pery|ene
Chrysene
Dibenz[a,/7]anthracene
Fluoranthene
Fluorene
lndeno[/,2,3-codpyrene
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 highly
perfused tissues, such as the lung, liver, gastrointestinal tract, and kidney. Animal
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5.6 POLYCYCLIC AROMATIC HYDROCARBONS (PAHs)
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, 1995).
PAHs have been shown 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
There is little data 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 ingestion, and the liver and the skin have been identified as target organs
in animals after oral or dermal exposure, respectively (ATSDR, 1995). Death has
been observed in animals after parenteral exposure to a number of PAHs (ATSDR,
1995). The intraperitoneal LD50 values in mice for pyrene, anthracene, and
benzo[a]pyrene are 514, >430, and 232 mg/kg, respectively.
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, 1995). Chronic studies in animals exposed
to PAHs by ingestion, intratracheal installation, or skin-painting have not identified
adverse health effects other than cancer.
5.6.5 Developmental Toxicity
No information is available regarding the developmental toxicity of PAHs in
humans. In vitro studies suggest that human placental endocrine and hormonal
function may be adversely affected by exposure to benzo[a]pyrene (ATSDR,
1995). Animal data describing developmental effects are mostly limited to
benzo[a]pyrene administered orally or parenterally and indicate that PAHs have
the potential to induce adverse developmental effects such as resorptions and
malformations, testicular changes including atrophy of the seminiferous tubules
and interstitial cell tumors, immunosuppression, and somatic tumor induction.
5.6.6 Mutagenicity
Benzo[a]pyrene has been thoroughly studied in genetic toxicology test systems
(ATSDR, 1995). 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, DNA adduct formation, unscheduled DNA synthesis, sister chromatid
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5.6 POLYCYCLIC AROMATIC HYDROCARBONS (PAffS)
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, 1995). 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,fc]anthracene). Dermal exposure of animals to
benz[a]anthracene, benzo[a]pyrene, benzo[/?]fluoranthene, benzo[/c|fluoranthene,
chrysene, dibenz[a,/7]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 currently verified by EPA's Carcinogenicity Risk
Assessment Verification Endeavor Work Group (IRIS, 1994) are listed below:
Acenaphthylene
Anthracene
Benz[a]anthracene
Benzo[a]pyrene
Benzo[6]fluoranthene
Benzo[/c]fluoranthene
Benzo[gr,/7,/]perylene
Chrysene
Dibenz[a,/?]anthracene
Fluoranthene
Fluorene
lndeno[ 7,2,3-co]pyrene
Phenathrene
Pyrene
D (not classifiable as a human carcinogen)
D
B2 (probable human carcinogen)
B2
B2
B2
D
B2
B2
D
D
B2
D
D .
The EPA and others have developed a relative potency estimate approach for the
PAHs (Nisbet and LaGoy, 1992; U.S. EPA, 1993s). 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.
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5.6 POLYCYCLIC AROMATIC HYDROCARBONS (PAHs)
Table 5-2. Toxicitv Equivalent Factors for Various PAHs
Compound
Toxicity Equivalency Factor (TEF)
Dibenz[a,/7]anthracene
Benzo[a]pyrene
Benz[a]anthracene
Benzo[Jb]fluoranthene
Benzo[/c]fluoranthene
Indeno[ 1,2,3-ccflpyrene
Anthracene
Benzo[g,/y]perylene
Chrysene
Acenaphthene
Acenaphthylene
Fluoranthene
Fluorene
Phenathrene
Pvrene
5
1
0
0.1
0.1
0.1
0.01
0.01
0.01.
0.001
0.001
0.001
0.001
6.001
o.oo-i
Source: Nisbet and LaGoy (1992).
U.S. EPA (1993s) has derived relative potency estimates based on mouse skin
carcinogenesis. These are shown in Table 5-3.
5.6.8 Special Susceptibilities
ATSDR has indicated 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 (ATSDR, 1995). 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, due to increased permeability of the embryonic and fetal blood-brain
barrier and the immaturity of the enzymatic systems that are responsible for
elimination.
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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
Benz[a]anth racene
Benzo[b]fluoranthene
Benzo[/c|fluoranthene
Chrysene
Dibenz[a,/7]anthracene
lndeno[ 1,2,3-ccflpyrene
1.0
0.145
0.167
0.020
0.0044
1.11
0.055b
Source: U.S. EPA, 1993s.
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.
5.6.9 Interactive Effects
Because humans are usually exposed to PAHs in complex mixtures rather than
to individual PAHs, it is important to understand the potential interactions between
the PAHs and other components of the mixture (ATSDR, 1995). 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 noncarcinogenic
and carcinogenic PAHs has been extensively examined in animals. Weakly
carcinogenic or noncarcinogenic PAHs, including benzo[e]pyrene, benzo[g,h,i\
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]fluorantherie, benzo[/^fluoranthene, chrysene, and a
mixture of anthracene, phenathracene, and pyrene have been shown to sig-
nificantly 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.
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
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5.6 POLYCYCLIC AROMATIC HYDROCARBONS (PAHs)
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.
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, 1995).
5.6.11 Summary of EPA Levels of Concern
Carcinogenicity (benzo[a]pyrene) 7.3 per mg/kg/d.
5.6.12 Major Sources
ATSDR (1995), IRIS (1997), U.S. EPA (1995).
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5.7 POLYCHLORINATED BfPHENYLS (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 Arocior, with a numeric designation that indicated their chlorine content.
Although production and use were banned in 1979, the chemical group is
extremely persistent in the environment and bioaccumulates through the food
chain. There is evidence that some dioxin-like PCB congeners, which are
assumed to be the most toxic, preferentially accumulated in higher organisms.
Consequently, the aggregate toxicity of a PCB mixture may increase as it moves
up the food chain (U.S. EPA, 1993a). As a result of this, the congeneric
composition of PCB mixtures found in fish tissue may differ significantly from the
environmental PCB source. Often the mixtures of interest are not those that have
been used in studies of laboratory animals to determine toxicity. The preferable
studies, under these conditions, are 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. When reliable human data are lacking, animal
data may need to be used.
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, 1995; U.S. EPA, 1993a). Due to the lack of congener-specific
information, the Office of Water recommends, as an interim measure, that total
PCB concentrations be reported as the sum of Aroclors. The first volume in this
document series, Sampling and Analysis, contains a detailed discussion of
analysis of this group of chemicals (U.S. EPA, 19953a).
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 oral absorption efficiency of 75 to greater than 90 percent
in rats, mice, and monkeys (IRIS, 1997). Due to their lipophilic nature, PCBs,
especially the highly chlorinated congeners, tend to accumulate in lipid-rich
tissues. Greater relative amounts of PCBs are usually found in the liver, adipose,
skin, and breast milk. Human milk may contain a large amount of PCBs due to
their high fat content (ATSDR, 1995). A Canadian study found human milk
concentrations more than 10 times higher than whole blood concentrations. This
is important because it has been shown that absorption of penta-, hexa-, and
heptachlorobiphenyls from breast milk by nursing infants may reach over 90
percent of the dose (ATSDR 1995). It has been estimated that, in some
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5.7 POLYCHLORINATED BIPHENYLS (PCBs)
industrialized countries, an infant may accumulate 6.8 percent of its lifetime PCB
body burden during an exclusive nursing period of 6 months (Kimbrough, 1995).
The PCB congener composition of milk differs from that of the PCB source.
Offspring can also be exposed to PCBs through placenta! transfer. PCBs have
also been measured in other body fluids including plasma, follicular fluid, and
sperm fluid. Indirect evidence of oral absorption in humans is available from
studies of subjects who consumed PCB-contaminated fish and PCB-contaminated
rice oil, from a volunteer who ingested a PCB mixture, and from nursing infants
(ATSDR, 1995). Pharmacokinetics data do not suggest route-specific target
organs.
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. Studies indicate that the
metabolism of PCBs by monkeys and rats is more similar to humans than other
species tested (IRIS, 1997). 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 clearance of congeners that are more easily metabolized
and retention of those that resist metabolism (ATSDR, 1995).
There are some data on the half-life of the various PCBs in humans. In a volunteer
who ingested a PCB mixture containing 54 percent chlorine, the elimination half-
lives from blood for two hexachlorobiphenyls and one heptachlorobiphenyl
congener ranged from 121 to 338 days (ATSDR, 1995). 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, 1995). In subjects who consumed PCB-contaminated rice in Taiwan, the
half-lives for several pentachlorobiphenyls ranged from 3 to 24 months.
PCBs induce mixed function oxidases and different congeners induce specific
forms (isozymes) of the cytochrome P-450 system. Although there has been much
research into the mechanisms of PCB toxicity, there has not been clear definition
of the mechanisms for most congeners. The congeners appear to act by a variety
of mechanisms (ATSDR, 1995). A few highly potent PCB congeners (dioxin-like
congeners) bind to a cytosolic protein, the Ah receptor, which regulates the
synthesis of a variety of proteins. The toxicity of these congeners is related to
steps that follow the initial binding with the Ah receptor. The toxicity of other PCB
congeners seems to be unrelated to the Ah receptor. Ultimately, the toxicity of a
PCB mixture may depend on the toxicity of the individual congeners and their
interactions. A detailed discussion of PCB pharmacokinetics is available in the
ATSDR Toxicological Profile for PCBs (ATSDR, 1995).
5.7.3 Acute Toxicity
Studies in animals have shown that exposure to very high PCB doses can cause
death. However, doses of such magnitude are unlikely in environmental exposures
and current industrial settings. There have been no reports of deaths in humans
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5.7 POLYCHLORINATED BIPHENYLS (PCBs)
after exposure to PCBs (ATSDR, 1995). Immature animals appear to be more
sensitive to acute lethal effects of PCBs than adults (ATSDR, 1995).
5.7.4 Chronic Toxicity
Numerous effects have been documented in studies in animals including hepatic,
gastrointestinal, hematological; dermal, body weight, endocrine, immunological!
neurological, reproductive, developmental, and liver cancer (ATSDR, 1995). Most
of the studies have involved oral exposure. Despite the variety of adverse effects
observed in animals exposed to PCBs, frank adverse effects in humans have not
' been observed. 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 (James et al.,1993; Kimbrough, 1995). There is also some
evidence suggesting that monkeys may be much more sensitive than humans.
EPA has derived an RfD of 2 x 10'5 mg/kg/d for Aroclor 1254 (IRIS, 1997). 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
in the monkeys, as well as significant reductions in antibody levels (IgM and IgG)
in response to injected sheep red blood cells at the lowest dose tested after
chronic treatment with Aroclor 1254. 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. This RfD is used to calculate
the consumption limits for noncarcinogenic effects for the general population listed
in Section 4.
EPA has medium confidence in the study used as the basis for the RfD, 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, 1997).
ATSDR has determined that immunological effects are a sensitive endpoint for
chronic toxicity and developed an MRL of 2 x 10'5 mg/kg/d based on such effects
(ATSDR, 1995). The studies used as the basis for the MRL are the same as those
listed above in the IRIS discussion. Uncertainty factors of 10 each for the use of
a LOAEL and for human variability were used. A factor of 3 was used for
extrapolation from animals to humans. Decreased IgG and IgM levels were noted.
Chronic toxicity in other organ systems (as listed above) was noted at exposure
levels higher than the LOAEL of 0.005 mg/kg/d (ATSDR, 1995).
5.7.5 Developmental Toxicity
PCB mixtures have been shown to cause adverse developmental effects in
experimental animals (ATSDR, 1995). Several studies in humans have also
suggested that PCB exposure may cause adverse effects in children and in
developing fetuses (U.S. EPA, 1995). However, study limitations, including lack of
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5.7 POLYCHLORINATED BIPHENYLS (PCBs)
control for confpunding variables, and deficiencies in the general areas of
exposure assessment, selection of exposed and control subjects, and the
comparability of exposed and control samples have obscured the interpretation of
the results (ATSDR, 1995).
The RfD for Aroclor 1016 is based on adverse developmental effects 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 trans-
placentally], 3 for interspecies extrapolation, 3 for database limitations [male
reproductive effects are not directly addressed in studies and two-generation
reproductive studies are not available], and 3 for extrapolation from subchronic to
chronic) to the NOAEL yields an RfD of 7 x ID'5 mg/kg/d (IRIS, 1997). 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 used as the basis for the RfD, in the
database, and in the RfD. 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, 1997).
The following discussion of developmental toxicity contains study information in the
following order: human data, short-term, intermediate length, and longer-term
studies, and a summary.
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). Although exposure quantification
was not precise, it has been estimated that the average exposure was 5 x 10"4
mg/kg/d. Contaminated fish consumption and levels of total PCBs in cord serum
correlated with lower birth weight, smaller head circumference, and shorter
gestational age. However, when the two populations were divided according to the
cord serum level, the great majority in the low-level group were fish eaters, which
suggested that fish consumption rates were poor indicators of PCB exposure. Fish
consumption, however, was correlated with delayed neuromuscular maturity, and,
at? months, the children had subnormal visual recognition memory. The exposure
estimates in this study were not precise and varied widely; the recall ranged over
a number of years with a mean consumption duration (as noted above) of 16 years
and the PCB concentrations in different types of fish of 168 ppb to 3,012 ppb.
Children from this cohort have been examined at age 4 and 11 years. At age 4,
cord serum PCB levels were associated with impaired short-term memory. Activity
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5.7 POLYCHLOR/NATED BIPHENYLS (PCBs)
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. Exposure during breast feeding, assessed based
on PCB concentrations in milk and the number of weeks of nursing, was not
associated with the results of the tests, neither was serum concentration of PCBs
at 11 years of age. The results from this series of studies were confounded by the
fact that there may have been maternal exposure to other chemicals and 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 (ATSDR, 1995).
A pharmacokinetics approach to estimating safe exposure levels has been taken
by the Great Lakes Sport Fish Advisory Task Force using the Michigan study data
(Anderson and Amrhein, 1993). This approach utilized relationships between milk
PCB concentrations, fish intake and concentrations, and developmental effects.
Assumptions were made regarding body weight (60 kg), percentage of body fat (25
percent), and the biological half-life of PCBs in humans (1 year) (Anderson and
Amrhein, 1993). The pharmacokinetics approach has the potential for introducing
more precision to the process of estimating thresholds and evaluating dose-
response relationships. However, it relies on the use of many physiological
variables, as well as dose and response values. In the specific case of the Great
Lakes approach, there is concern that the assumptions that were made for the
"average" women and "average" body fat composition do not take into
consideration the 49 percent of women who have above-average values. Although
this variability would introduce minimal alterations at values near the average,
there could be significant deviation from predicted values at the 75th or 90th
percentiles. The approach also assumes that reproduction occurs at 25 years of
age with the estimated body burden based on this assumption (Anderson and
Amrhein, 1993). Maternity over the age of 25 would entail greater exposure to the
fetus due to the higher maternal body burden associated with a longer
accumulation period.
A study of children born to women with background body burdens of PCBs in
North Carolina found no correlation between birth weight or head circumference
with PCB levels. The authors reported that neurobehavioral deficits observed
through 2 years of age were not detectable at ages 3, 4, and 5, based on
intellectual and motor function assays. Exposure was confounded by the presence
of DDE in b!6od and milk samples from the mothers, although it was shown that
some of the behavioral deficits were more closely associated with PCB exposure.
This study utilized PCB body burdens rather than intake as the measure of
exposure (ATSDR, 1995).
Four additional relevant studies were summarized by ATSDR (1995). A study of
women from the Green Bay, Wisconsin, area found no significant differences
between a control group and fish eaters regarding stillbirths, multiple births,
congenital anomalies, and low birth weight. Another study of PCB-contaminated
Lake Ontario sports fish found no consistent relationship between sports fish
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5.7 POLYCHLORINATED BIPHENYLS (PCBs)
consumption and/or PCB exposure and incidence of spontaneous fetal death.
However, since fetal death was the outcome measured, the results could not rule
out an effect of PCBs on reproduction. A third study compared a small number of
women who had spontaneous preterm delivery and a matched control group and
found no association between serum PCB levels and spontaneous preterm birth.
The fourth study examined infants born to mothers occupationally exposed to
PCBs. Infants born to mothers with high exposure had lower mean birth weight
and shorter mean gestational age than those born to low-exposure workers. After
adjustment for relevant covariates, it was concluded that the decreased birth
weight may have been mediated by exposure to high levels of PCBs. The authors
further indicated that the small difference in birth weight had no clinical significance
for term infants.
The results of animal studies generally support those observed in humans. Short-
term studies in animals exposed prenatally to PCBs have identified the following
effects: hydronephrosis in mice after a single dose of 244 mg/kg (Aroclor 1254) on
gestation day 9; no effects in mice following 12 daily doses of up to 12.5 mg/kg/d
(Aroclor 1254) on gestation days 6 to 18; fetal weight reduction in rats with 9 days
of dosing at 5 mg/kg/d with reduced survival at 15 mg/kg/d and a NOAEL of 2.5
mg/kg/d (Aroclor 1254); and impaired learning in rats at 4 mg/kg/d with 10 days of
dosing (Fenclor 42). Decreased survival was observed at higher doses. In
addition, decreased fertility was observed in male offspring of rats treated with ;>8
mg/kg/d (Aroclor 1254) during lactation. Based on the results of short-term
exposure assays, 'ATSDR concluded that neurobehavioral endpoints may be the
most sensitive for assessing developmental effects.
Intermediate-length exposure studies (e.g., during the prenatal and lactational
periods) indicate neurological, thyroid, liver, growth, and hormonal abnormalities
in offspring and reduced litter size (ATSDR, 1995). Delayed growth and 89 percent
neonatal death was reported in mink at 0.18 mg/kg/d (Aroclor 1254) and, therefore,
this exposure level constitutes an PEL (frank effect level). Fetal death was also
observed in monkeys following maternal treatment with 0.1 mg/kg/d Aroclor 1254.
Mink and monkeys appear to be more sensitive species for PCB-induced
developmental toxicity than rodents (ATSDR, 1995).
Information on chronic developmental toxicity is available from studies in monkeys
(ATSDR 1995). Exposure periods ranged from 12 to 37 months. The lowest
LOAEL was 0.005 mg/kg/d for inflammation of tarsal glands, nail lesions, and gum
recession in offspring of monkeys exposed to Aroclor 1254. Adverse
neurobehavioral effects were reported at 0.03 mg/kg/d for Aroclor 1016 and at 0.08
mg/kg/d for Aroclor 1248; the respective NOAELs were 0.007 and 0.03 mg/kg/d.
Other effects observed included reduction in birth weight (0.03 to 0.08 mg/kg/d)
and increased infant death with doses as low as 0.1 mg/kg/d for Aroclor 1248.
As mentioned above, exposure via lactation is a significant concern for neonates.
Animal studies indicate that lactational exposure may be more significant than
prenatal exposure. In monkeys, signs of PCB intoxication were observed in
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5.7 POLYCHLORtNATEP BfPHENYLS (PCBsJ
lactationally exposed offspring, but not in offspring exposed only prenatallv
(ATSDR, 1995).
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.
5.7.6 Mutagenicity
IRIS reports that the majority of mutagenicity assays of PCBs have been negative
(IRIS, 1997, for PCB mixtures).
An increase in the percentage of chromosomal aberrations in peripheral
lymphocytes was reported in a study of workers manufacturing PCBs for 10 to 25
years. Increased sister-chromatid exchange was also reported in that study.
Although workers and controls were matched for smoking and.drinking, concurrent
exposure to other known human genotoxic chemicals occurred (ATSDR, 1995).
A different study found increased incidence of chromatid exchanges in
lymphocytes from workers exposed to PCBs in an electric station fire compared
to unexposed controls. The possibility that toxic chlorinated dioxins and/or furans
generated during the fire may have been responsible for the effects could not be
ruled out.
ATSDR reports that most in vitro assays and in vivo animal assays yielded
negative results, although both positive and negative results were reported.
Positive study results include an increase in unscheduled DNA synthesis (ATSDR
1995). See also Section 5.7.9.
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. Recent revaluation of the animal data showed that PCBs
with 60 percent chlorine content consistently induced a high yield of liver tumors
in rats and that PCB mixtures with 54 or 42 percent chlorination have a lower
carcinogenic potential than those with 60 percent chlorine. Human epidemiological
studies of PCBs have not yielded conclusive results (Silberhorn et al.,1990). As
with all epidemiological studies, it is very difficult to obtain clear unequivocal results
due to 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 (IRIS, 1997).
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, 1996f), 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
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5.7 POLYCHLORINATED BIPHENYLS (PCBs)
range of cancer potency factors, then uses information on environmental
processes to provide guidance on choosing an appropriate potency factor for
representative classes of environmental mixtures and different pathways.
Depending on the specific applicaiton, either central estimates or upper bounds
can be appropriate. Central estimates describe a typical individual's risk, while
upper bounds provide assurance (i.e., 95 percent confidence) that this risk is not
likely to be underestimated if the underlying model is correct. 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 were 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 potency factors (using
a body weight ratio to the 3/4 power) ranging from 0.07 (lowest risk and
persistence) to 2 (high risk and persistence) per mg/kg/d for average lifetime
exposures to PCBs. It is noteworthy that bioaccumulated PCBs appear to be more
toxic than commercial PCBs and appear to be more persistent in the body. For
exposure through the food chain, risks can be higher than other expsosures.
The high risk and persistence cancer slope factor of 2.0 per mg/kg/d was used to
calculate the carcinogenicity fish consumption limits, because the major pathway
of exposure to persistent toxic substances such as PCBs is via dietary exposure
(i.e., contaminated fish consumption).
5.7.8 Special Susceptibilities
ATSDR has indicated that embryos, fetuses, and neonates are unusually
susceptible to PCBs due to their underdeveloped enzymatic systems, which may
cause delayed elimination and, therefore, accumulation of PCBs in the body.
Breast-fed infants are at particular risk because a steroid secreted in human milk,
but not cows' milk, inhibits glucuronyl transferase activity, which is critical to PCB
metabolism and excretion (ATSDR, 1995).
Other individuals at potentially greater risk include those with syndromes
associated with incompletely developed glucuronide conjugation mechanisms,
those with hepatic infections, compromised liver functions, or acute intermittent
porphyria (ATSDR, 1995).
In addition, PCBs cause induction of the mixed function oxidase system.
Individuals exposed to chemicals (including Pharmaceuticals) that rely on the
mixed function oxidase system for activation or detoxification may experience
altered effectiveness of the chemicals. Further discussion may be found in
Appendix C under "Organochlorines."
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5.7 POLYCHLOR1NATED BIPHENYLS (PCBs)
5.7.9 Interactive Effects
PCBs induce microsomal enzymes. See Appendix C under "Organochlorines" for
potential interactions arising from this characteristic.
ATSDR reports that:
The genotoxicity of numerous carcinogens is potentiated in vitro by
PCBs, but this does not indicate that PCBs should be regarded
universally as tumor promoters because of the protective role of
PCBs against carcinogenicity of many genotoxic carcinogens in
vivo (ATSDR, 1995).
MIXTOX reports potentiation between PCBs and mMirex in a rat dietary study.
Other studies of this combination have not found interactive results (MIXTOX).
5.7.10 Critical Data Gaps
A joint team of scientists from EPA, ATSDR, and NTP have identified the following
data gaps: human epidemiological studies; genotoxicity studies of various mixtures
of PCBs including cytogenetic analysis of human populations exposed to PCBs;
reproduction studies in humans and animals including fertility studies in males of
a sensitive species; developmental studies including histological examination of
developing neurological tissues in experimental animals, neurodevelopmental
studies designed to identify NOAELs, and immunological studies in animals
exposed in utero; 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; human studies on
PCBs and hypertension and liver toxicity; pharmacokinetics studies; and studies
to elucidate the differing toxicities of the various congeners comprising PCB
mixtures; studies to elucidate the mechanisms and significance of Ah-receptor-
independent effects (ATSDR, 1995).
5.7.11 Summary of EPA Levels of Concern
Developmental Toxicity
Chronic Toxicity
Carcinogenicity
2 x 10"5 mg/kg/d based on Aroclor 1254
2 x 10"5 mg/kg/d based on Aroclor 1254
2.0 per mg/kg/d based on mixed PCBs.
5.7.12 Major Sources
ATSDR (1995), HSDB (1993), IRIS (1997), James et al. (1993), Kimbrough (1995)
Silberhorn et al. (1990), U.S. EPA (1996f).
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5.8 DIOXINS
5.8 DIOXINS
5.8.1 Background
Dioxin has been undergoing extensive review within EPA for several years.
Consequently, only a brief summary, taken from Volume 1 of this guidance series
(second edition), 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 Docu-
ment for2,3,7,8-Tetrachlorodibenzo-p-dioxin (TCDD) and Related Compounds
(three volumes) (U.S. EPA, 1994b) is undergoing final review. It is anticipated that
the dioxin reassessment document will be sent for final external peer review during
the summer of 1997. Following peer review, the document will be sent to the
Agency's Science Advisory Board for final review by the fall of 1997. The final
dioxin reassessment document is scheduled for release in 1998.
Dioxin is a generic term that is used, in this case, to specify 2,3,7,8-tetra-
chlorodibenzo-p-dioxin (TCDD). 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. Alternatively, 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).
Dioxin is extremely toxic to humans and animals and affects multiple organ
systems. Adverse effects observed in animal studies include teratogenicity,
fetotoxicity, reproductive dysfunction, carcinogenicity, and immunotoxicity (U.S.
EPA, 1993a). Dioxin has the highest cancer potency in animals of the chemicals
evaluated by EPA. A cancer-risk-based health advisory can be calculated using
the existing cancer slope factor of 1.56 x 10+5 per mg/kg/d (U.S. EPA, 1993a).
5.8.2 Summary of EPA Levels of Concern
Carcinogenicity 1.56 x 10+5 per mg/kg/d.
5.8.3 Major Source
U.S. EPA(1993a).
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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
communication. 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
and can be used 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
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6. MAPPING TOOLS
concentrations can be combined to produce an overview of populations that may
be at risk. Further discussion of mapping as a technique for risk communication
is included in Volume 4 of this guidance series, Risk Communication. 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).
6.3 BASIC GIS 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 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 a Superfund
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 dis-
advantages 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 precisely located so that different data layers will overlay
correctly. Geographic information contains either an explicit geographic reference
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6. MAPPING TOOLS
ft «,.
WoBd
Figure 6-1. GIS Data Layers May Use Raster or
Vector Representation Techniques.
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 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
polygon 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
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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).
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 and the World Wide Web
(WWW or WEB) are making it easier to locate and obtain, often for free), a variety
6-4
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6. MAPPING TOOLS
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. 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 INTERNET SOURCES OF EXISTING DATA FILES AND GIS 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 private firms
called the National Spatial Data Infrastructure (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 tasked 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
by subscribing with companies known as Internet Service Providers. Internet
access is also available through colleges, libraries, research institutes, and
government agencies. Web sites are identified by special addresses called
Universal Resource Locators (URLs). The URL providing general information for
the entire National Spatial Data Infrastructure is:
http://fgdc.er.usgs.gov/
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 United States Department of
Agriculture (USDA), and EPA have their own NSDI WEB pages with links to more
specialized data items. EPA's link to the NSDI is at
http://nsdi.epa.gov/nsdi/
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 agency programs. Examples include a facility called SURF YOUR
WATERSHED that 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
6-5
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6. MAPPING TOOLS
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 Toxics Release
Inventory (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 new Web tool
called BASINInfo that can produce displays of the major types of permitted
facilities within a USGS Cataloging Unit.
EPA's SURF YOUR WATERSHED facility provides an on-line set of maps derived
from the Office of Science and Technology's North American Fish and Wildlife
Consumption Database (NAFWCD). Figure 6-3 shows a display depicting the
• locations of active advisories for the State of North Carolina. GIS maps showing
the location of fish advisories in any of the 50 States, U.S. Territories, and the
District of Colombia 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
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
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6. MAPPING TOOLS
V
Figure 6-2. Examples of GIS Displays from EPA's
BASINInfo Maps-on-Demand Facility.
Figure 6-3. Map Showing Active Fish
and Wildlife Advisories for a State.
6-7
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6. MAPPING TOOLS
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.
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. 1DRISI (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 program cost was $650 (government rate) in 1995. 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
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6. MAPPING TOOLS
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 provided
to EPA, which are widely applicable, are especially welcome.
6-9
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7. LITERATURE CITED
SECTION 7
LITERATURE CITED1 2
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The addresses from which to obtain State documents are listed in Appendix A, Sources of Additional
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2 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
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polychlorinated biphenyl residues in human breast lipids and their relation to
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Switzerland: WHO.
Wulf, H.C., N. Kromann, N. Kousgaard, et al. 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 Environment 48:81 -94.
Young, Pat. 1994. U.S. Environmental Protection Agency Region 9. Personal
communication with Abt Associates, July 28.
Zabik, M.E., et al. 1993. Assessment of Contaminants in Five Species of Great
Lakes Fish at the Dinner Table. Final Report to the Great Lakes Protection
Fund, March.
7-19
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APPENDIX A
SOURCES OF ADDITIONAL INFORMATION
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APPENDIX A
APPENDIX A
SOURCES OF ADDITIONAL INFORMATION
The sources listed below were consulted in the development of this document and
may be useful to readers who wish to obtain more in-depth information on specific
topics. A listing of the State documents and addresses begins on page A-5.
ATSDR (Agency for Toxic Substances and Disease Registry). U.S. DHHS, PHS.
See Toxicological Profiles for Various Chemicals. Obtain from ATSDR in
Atlanta, GA.
Baker, S.R., and C.F. Wilkinson (eds). 1988. The Effects of Pesticides on Human
Health: Advances in Modern Environmental Toxicology Volume 18.
Princeton, NJ: Princeton Scientific Publishing.
Crump, K.S. 1981. Statistical aspects of linear extrapolation. In: Proceedings of the
3rd Life Sciences Symposium, Health Risk Analysis. Richmond, C., and E.
Copenhaver (eds.). Philadelphia, PA: Franklin Institute Press.
Hood, R.D., Ed. 1989. Developmental Toxicology: Risk Assessment and Future.
For: Reproductive and Developmental Toxicology Branch, OHEA, USEPA.
New York: Van Nostrand Reinhold.
HSDB (Hazardous Substances Data Bank). 1992. On-line from Toxnet.
Humphrey, H.E.B. 1988. Chemical contaminants in the Great Lakes: the human
health aspect. In: M.S. Evans (ed.). Toxic Contaminants and Ecosystem
Health: A Great Lakes Perspective. New York: John Wiley and Sons, p. 153-
165.
IARC (International Agency For Research on Cancer). 1986. Multiple volumes on
a variety of chemical contaminants. Obtain from World Health Organization,
IARC, Lyon, France.
IRIS (Integrated Risk Information System). 1993. Chemical-specific and
background document files are available. Obtain through Toxnet. Developed
by EPA.
Klaassen, C.D., M.O. Amdur, and J. Doull (eds.). 1986. Casarett and Doull's
Toxicology: The Basic Science of Poisons. New York: MacMillan Publishing
Company.
A-3
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APPENDIX A
Kuehl, D.W., B.C. Butterworth, A. McBride, S. Kroner, and D. Bahnick. 1989.
Contamination of fish by 2,3,7,8-tetrachlorodibenzo-p-dioxin: A survey of fish
from major watersheds in the United States. Chemosphere 18:1997-2014.
Review of Environmental Contamination and Toxicology. 1988. USEPA, Off ice of
• Drinking Water Health Advisories, (includes multiple pesticides), Vol. 104,
New York: Springer-Verlag.
TVA (Tennessee Valley Authority). 1992. Use of Risk Assessment Techniques to
Evaluate TVA's Fish Tissue Contaminant Data, Tennessee Valley Authority
Water Resources Division (prepared by Janice P. Cox).
U.S. EPA (Environmental Protection Agency). Memoranda titled "List of Chemicals
Evaluated for Carcinogenic Potential" issued regularly by the Office of
Pesticide Programs.
. Tox One Liners for Pesticides of Interest. Obtain from Office of
Pesticide Programs, Washington, DC.
. 1988. Proposed guidelines for assessing female reproductive risk. U.S.
EPA. Federal Register, 53: 24834-24847. [Cited in U.S. EPA, 1992.]
_. 1988. Proposed guidelines for assessing male reproductive risk. U.S.
EPA. Federal Register, 53: 24850-24869. [Cited in U.S. EPA, 1992.]
. 1989. Interim Methods for Development of Inhalation Reference Doses.
Office of Health and Environmental Assessment, Washington, DC.
_. 1989. Workshop Report on EPA Guidelines for Carcinogen Risk
Assessment. Risk Assessment Forum, Washington, DC.
. 1989. Workshop Report on EPA Guidelines for Carcinogen Risk
Assessment: Use of Human Evidence, Risk Assessment Forum,
Washington, DC.
. 1991. Risk assessment guidance for Superfund. Volume 1: human
health evaluation manual supplemental guidance—"standard default
exposure factors." Office of Emergency and Remedial Response.
. . Office of Pesticide Programs RfD Tracking Report—obtain the most
recent version. It is usually issued quarterly.
. 1993. A SAB Report: Cholinesterase Inhibition and Risk Assessment,
Review of the Risk Assessment Forum's Draft Guidance on the Use of Data
on Cholinesterase Inhibition in Risk Assessment by the SAB/SAP Joint
Committee, Washington, DC-
A-4
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APPENDIX A
U.S. EPA (Environmental Protection Agency). 1993. Review of the Methodology
for Developing Ambient Water Quality Criteria for the Protection of Human
Health. Prepared by the Drinking Water Committee of the Science Advisory
Board, Washington, DC.
_. 1996. Listing of Fish and Wildlife Advisories. EPA-823-C-97-004 (3.5-
inch diskettes) or EPA-823-C-97-005 (CD-ROM). Cincinnati!, OH: National
Center for Environmental Publications and Information. Also available from
the Internet at: http://www.epa.gov/ost.
U.S. FDA (Food and Drug Administration). Guidance Documents for Chemicals in
Fish and Shellfish. May be obtained from the Center for Food Safety and
Applied Nutrition, U.S. FDA, Washington, DC.
STATE DOCUMENTS
The documents are listed in alphabetical order by State. The address, which
follows the first document listed for each State, is the location of the office(s)
that have contributed to this document.
Monterey Bay Marine Environmental Health Survey: Health Evaluation, California
Environmental Protection Agency, Office of Environmental Health Hazard
Assessment, 1992.
California Environmental Protection Agency
Office of Environmental Health Assessment
Pesticide and Environmental Toxicology Section
601 N. 7th Street, P.O. Box 942732
Sacramento, CA 94234-7320
A Study of Chemical Contamination of Marine Fish From Southern California II.
Comprehensive Study, California Environmental Protection Agency, Office
of Environmental Health Hazard Assessment, 1991.
California Sport Fishing Regulations, California (State of) Fish and Game
Commission, Department of Fish and Game, 1992.
Exposure Guidelines for Mercury in Fish: Memorandum to Charles S. Mahan, M.D.
from Richard W. Freeman, Ph.D., Florida (State of) Department of Health
and Rehabilitative Services, 1993.
Florida Department of Health and Rehabilitative Services
Toxicology and Hazard Assessment
1317 Winewood Blvd.
Tallahassee, FL 32399-0700
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APPENDIX A
Advisories on Mercury Concentrations in Large Mouth Bass: memorandum to MRS
District Administrators from Charles S. Mahan, M.D., Florida (State of)
Department of Health and Rehabilitative Services, 1993.
Guidelines for Issuing Advisories/Bans on the Consumption of Chemically
Contaminated Fish, Louisiana Department of Health and Hospitals, Office
of Public Health, Environmental Epidemiology Section, 1990 (updated
1991).
Louisiana Department of Health and Hospitals
Office of Public Health
New Orleans, LA 70160
Summary of Revisions to the Michigan Sport Fish Consumption Advisory,
Michigan Department of Public Health, Division of Health Risk Assessment,
1992.
Michigan Department of Public Health
Division of Health Risk Assessment
3423 N. Martin Luther King Blvd.
Lansing, Ml 48909
Health Risk Assessment for the Consumption of Sport Fish Contaminated with
Mercury, PCBs, and TCDD, Minnesota Department of Health (Pamela
Shubat), 1993.
Minnesota Department of Health
Division of Environmental Health
Section of Health Risk Assessment
925 S.E. Delaware Street, P.O. Box 59040
Minneapolis, MN 55459-0040
PCB Reference Dose for Effects on Fetal Development: Development of a
reference dose based on human exposure data, Minnesota Department of
Health (Pamela Shubat), 1992.
Fish Facts: Methylmercury in Fish, Minnesota Department of Health, 1991.
Health Risk Assessment for the Consumption of Sport Fish Contaminated with
Mercury, PCBs and TCDD: Guidelines for the 1991-1992 Minnesota Fish
Consumption Advisory, Minnesota Department of Health (Pamela Shubat),
1991.
Criteria used to issue fish consumption advice: 1992 Minnesota Fish Consumption
Advisory, Minnesota Department of Health (Pamela Shubat), 1992.
Fish Facts: Contaminants in Lake Superior Fish, Minnesota Department of Health,
1991.
A-6
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APPENDIX A
Fish Facts: Eating Minnesota Fish: Health risks and benefits, Minnesota Depart-
ment of Health, 1991.
Which Fish Are Safe To Eat?, Minnesota Department of Health.
Eating Minnesota Fish: A Guide to Your Health, Minnesota Department of Health,
1991.
Minnesota Fish Consumption Advisory, Minnesota Department of Health, 1992
Fish Tissue Criteria for Dioxin, Mississippi Department of Environmental Quality,
1990.
Mississippi Bureau of Pollution Control
121 Fairmont Plaza
Pearl, MS 39208
Principles for Determining When and Where to Issue Health Advisories on
Chemical Contamination in Fish, Missouri Department of Conservation Fish
Contaminant Project, 1989.
Missouri Department of Health
Bureau of Environmental Epidemiology
P.O. Box 570
1730 E. Elm St.
Jefferson City, MO 65102
State Health Officials Issue New Advisories for Chlordane-Contaminated Fish on
Rivers, Lakes, Missouri Department of Health, 1991.
Missouri Department of Health 1992 Fish Consumption Advisory, Missouri
Department of Health, 1992.
A Study of Dioxin (2,3,7,8-Tetrachlorodibenzo-p-Dioxin) Contamination in Select
Finfish, Crustaceans and Sediments of New Jersey Waterways, New Jersey
Department of Environmental Protection, Office of Science and Research
(Thomas J. Belton, M.A.; Robert Hazen, Ph.D.; Bruce E. Ruppell; Keith
Lockwood; Robert Mueller, M.S.; Edward Stevenson; JoAnn J. Post, M.A.),
1985.
New Jersey Department of Environmental Protection
Division of Science & Research
CN 409
401 E. State St., 41E
Trenton, NJ 08625
A-7
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APPENDIX A
Environmental Impact Statement for Policy on Contaminants in Fish (Final), New
York State Department of Environmental Conservation, 1985.
New York State Department of Environmental Conservation
50 Wolf Rd.
Albany, NY 12233
The Procedure for North Dakota's Public Health Advisory Regarding Consumption
of Fish Contaminated with Methylmercury, North Dakota State Department
of Health & Consolidated Laboratories, Environmental Health Section
(Martin R. Schock, Special Studies Coordinator; Francis J. Schwindt, Chief),
1991.
North Dakota Department of Health and Consolidated Laboratories
Environmental Health Section
1200 Missouri Avenue
P.O. Box 5520
Bismarck, ND 58502-5520
A-8
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APPENDIX B
MUTAGEN1CITY AND GENOTOXICITY GUIDELINES
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APPENDIX B
APPENDIX B
MUTAGENICITY AND GENOTOXICITY GUIDELINES
The U.S. Environmental Protection Agency (EPA) has developed and published
Guidelines for Mutagenicity Risk Assessment (U.S. EPA, 1986). The information
in this appendix was primarily taken from that document. Quantitative assessment
of mutagenicity requires two steps: (1) determining the heritable effect per unit of
exposure (dose-response) and the relationship between mutation rate and disease
incidence, and (2) combining dose-response information with anticipated levels
and patterns of human exposure in order to derive a quantitative assessment of
risk (U.S. EPA, 1986). Current EPA guidance on mutagenicity risk assessment
specifies that:
Dose-response assessments can presently only be performed using
data from in vivo, heritable mammalian germ-cell tests, until such time
as other approaches can be demonstrated to have equivalent pre-
dictability. (U.S. EPA, 1986).
The relationship between in vitro assay results and effects in mammalian systems
is not sufficiently characterized to be able to use in vitro assays as the basis for
developing a dose-response assessment.
An example of the type of study that could be used for mutagenicity risk
assessment is an assay that directly detects heritable health effects in the first-
generation offspring. Human risk estimates are obtained by extrapolating the
induced mutation frequency or observed phenotypic effect downward to the
anticipated level of human exposure (EPA, 1986). No one extrapolation model has
been identified as the most appropriate. The Agency notes that departures from
linearity at low exposure and exposure rates has been observed for at least one
chemical. According to EPA, "[t]he Agency will consider all relevant models for
gene and chromosomal mutations in performing low-dose extrapolations and will
choose the most appropriate model. This choice will be consistent with both the
experimental data available and current knowledge of relevant mutational
mechanisms" (U.S. EPA, 1986).
The factors that should be considered in evaluating chemicals for mutagenic
activity include:
B-3
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APPENDIX B
• Genetic endpoints -, . ,
• Sensitivity and predictive value of the test systems for various classes of
chemical compounds
• Number of different test systems used for detecting each genetic endpoint
• Consistency of the results obtained in different test systems and different
species
• Aspects of the dose-response relationship • . -
• Whether the tests are conducted in accordance with appropriate test protocols
agreed upon by experts in the field (U.S. EPA, 1986).
Although there are often no in vivo data available on a chemical, there are in vitro
assay results for most common chemicals. These data may be used qualitatively
to evaluate the mutagenicity of chemicals. They are often used as supporting
evidence in carcinogenicity, developmental toxicity, and reproductive toxicity
evaluations.
Various types of test results may be obtained regarding mutagenicity. In evaluating
interactions in the mammalian gonad, two possible types of evidence have been
specified. Evidence for chemical interactions in the mammalian gonad may be
considered sufficient if it is demonstrated that "an agent interacts with germ-cell
DNA or other chromatin constituents, or that it induces endpoints such as
unscheduled DNA synthesis, sister-chromatid exchange, or chromosomal aberra-
tions in germinal cells" (U.S. EPA, 1986). Suggestive evidence of interaction in the
mammalian gonad "includes effects such as sperm abnormalities following acute,
subchronic or chronic toxicity testing, or finding of adverse reproductive effects
such as decreased fertility, which are consistent with the chemical's interaction
with germ cells" (U.S. EPA, 1986).
In practice, the outcomes of developmental and reproductivity toxicity testing often
do not indicate the type of toxicity that leads to effects such as decreased fertility.
The causes of decreased fertility range from mutagenicity leading to early fetal
death or failure to implant to maternal, paternal, or fetal toxicity. Positive significant
mutagenicity studies along with fetal toxicity or reduced fertility may be suggestive
that mutagenic action was a causal action. Developmental toxicity has been
studied in most chemicals (other than frank teratogens) relatively recently. The use
of the data from these studies with other types of data such as mutagenic,
pharmacokinetic, and reproductive system studies is developing; however,
currently there is not clear guidance on these types of evaluations.
EPA has addressed the issue of weight-of-evidence for mutagenicity by providing
a classification scheme with categories presented in decreasing order of strength
of evidence:
B-4
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APPENDIX B
1. Positive data derived from human germ-cell mutagenicity studies
2. Valid positive results from studies on heritable mutational events in mammalian
germ cells
3. Valid positive results from mammalian germ-cell chromosome aberrations
studies that do not include an intergenerational test
Reference
4. Sufficient evidence for a chemical's interaction.
U.S. EPA (Environmental Protection Agency). 1986. Guidelines for mutagenicity
risk assessment. Federal Register. 51(185):34006-34012.
B-5
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APPENDIX C
TOXICITY CHARACTERISTICS OF GROUPS OF ANALYTES
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APPENDIX C
APPENDIX C
TOXICITY CHARACTERISTICS OF GROUPS OF ANALYTES
Many chemicals on the list of target analytes (see Table 1-1 in Section 1) fall into
two groups: organochlorine pesticides and organophosphate pesticides.
Chemicals in these groups, while having many individual characteristics, have
multiple toxicity attributes in common with other members of the same group due
to their structural similarity. Rather than listing all aspects of the toxicity of each
member under the individual chemical discussions, those characteristics shared
by members of the group are listed in this appendix.
The following information is of a qualitative nature and includes the spectrum from
acute high-dose responses to chronic exposure low-dose responses. This range
was included for a number of reasons. Individual responses to chemical exposures
will vary considerably. It is not anticipated that all people exposed at the same
dose will respond in the same manner. Those individuals who are chemically
sensitive may respond to chronic low doses as severely as less sensitive
individuals who are exposed to high doses. Second, the effects elicited by
organophosphate and organochlorine pesticides can be characterized on a
continuum for many organ systems (e.g., nervous system effects, liver damage).
Therefore, many effects are associated with both acute and chronic exposure.
Finally, there are very limited dose-response data establishing specific human
thresholds for effects to occur. The risk values (reference doses), derived primarily
from animal studies, do not predict the exposure level at which response will occur,
but rather incorporate uncertainty factors with the study data to determine a level
at which no one is anticipated to experience adverse effects. Consequently, it is
not possible, in most cases, to provide quantitative data regarding the dose
associated with a specific level of effect in a specific organ system. Under most
circumstances it is assumed that the very severe effects such as convulsions,
coma, and death are associated with acute high-level exposures to chemicals. The
information provided below was obtained from the following sources (full citations
are provided at the end of this appendix):
• Recognition and Management of Pesticide Poisonings (U.S. EPA, 1982)
• Casarett and Doull's Toxicology (Klaassen et al., 1986)
• Pesticides and Human Health (Cunningham-Burns and Hallenbeck, 1984)
• Pesticides Studied in Man (Hayes, 1982)
• Handbook of Pesticide Toxicology (Hayes and Laws, 1991).
In addition, the Hazardous Substances Data Bank (HSDB, 1993), which was
consulted for specific information on target analytes, contains general clinical
C-3
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APPENDIX C
effects information for organophosphate exposures and organochlorine exposures
that was used in the development of this appendix.
C.1 Organochlorine Pesticides
Organochlorine pesticides are readily absorbed via the digestive system. They
often accumulate in fatty tissue, including brain and adipose tissue, and may also
be found in human milk due to its high lipid content. The neurological effects of
organochlorine exposure are based upon interference with axonic transmission of
nerve impulses. This causes altered functioning of the nervous system, primarily
the brain.
The following symptoms are commonly associated with exposure to organo-
chlorines: behavioral changes, sensory and equilibrium disturbances, involuntary
muscle activity, depression of vital centers (particularly those controlling respira-
tion), myocardial irritability, tremor, twitching, nausea, confusion, apprehension,
excitability, dizziness, headache, disorientation, weakness, paresthesias, con-
vulsions, and unconsciousness (HSDB, 1993; U.S. EPA, 1982).
Organochlorines stimulate synthesis of hepatic drug-metabolizing microsomal
enzymes, primarily in the liver; however, they do so in different ways (Hayes and
Laws, 1991, p. 739). Many organochlorines are associated with liver and kidney
toxicity. Organochlorine pesticides' induction of the hepatic microsomal enzyme
system causes alterations in the rate of metabolism of all other endogenous or
exogenous chemicals metabolized by this system. Metabolism may detoxify or
increase the toxicity of chemicals, depending on whether the parent or metabolite
is more toxic. For example, DDT is reported to promote some tumorigenic agents
and antagonize others as a result of the induction of microsomal enzymes
(ATSDR, 1992). In addition, exposure to other chemicals that induce the same
enzymes may increase the toxicity of the chemical under evaluation by enhancing
its metabolism to its toxic intermediate.
The induction of microsomal enzymes by organochlorines has serious implications
for the metabolism of some pharmaceutical drugs. Alterations in response to drugs
have been observed both in humans and experimental animals. 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 (HSDB, 1993). Concern regarding interaction with drugs is
indicated in discussions of especially susceptible populations in the Agency for
Toxic Substances and Disease Registry (ATSDR) Toxicological Profiles. For
example, due to the interactive effects of chlordane with other chemicals via
microsomal enzymes, ATSDR has cautioned that: "doses of therapeutic drugs and
hormones may require adjustment in patients exposed to chlordane" (ATSDR,
1992a). A similar caution is provided for DDT and its analogs: individuals who use
medications that involve the mixed function oxidase system (MFO inhibitors)
directly or for metabolic processes may be at risk for alteration of the drugs'
efficacy and/or timing if they are exposed to DDT (ATSDR, 1992b).
C-4
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APPENDIX C
For most chemicals, information was not available on the quantitative relationships
between various Pharmaceuticals and organochlorine body burdens or intakes.
When information was located on these types of interactions, it was included in the
chemical discussions. For more information on this, the reader is referred to the
ATSDR Toxicological Profiles and to the open literature.
C.2 Organophosphate Pesticides
Organophosphates are efficiently absorbed via ingestion. Their toxicity depends
in part on the rate at which the chemicals are metabolized in the body. This occurs
principally by hydrolysis to nontoxic or minimally toxic byproducts. The most
studied and obvious effect of organophosphate poisoning is cholinesterase
inhibition. This results from phosphorylation of the acetylcholinesterase enzyme
at nerve endings. Loss of enzyme function allows accumulation of acetylcholine
(the neurotransmitter) at cholinergic neuroeffector junctions, causing muscarinic
effects, and at skeletal myoneural junctions, and in autonomic ganglia (nicotinic
effects). Organophosphates cause central nervous system (CNS) disturbances
through impairments of nerve impulse transmission in the brain (U.S. EPA, 1982).
It is not clear what, if any, adverse effects are associated with exposure at levels
that produce only cholinesterase inhibition in the absence of any other effects. This
is currently under evaluation at EPA. In 1993, EPA's Scientific Advisory Board
(SAB) issued a report on cholinesterase inhibition and risk assessment (U.S. EPA,
1993). A key finding was:
To date, analyses of studies of cholinesterase inhibition in plasma and
in red blood cells do not provide information useful for evaluating
potential hazards and risks in the nervous system. This finding justifies
a new science policy against the use of blood cholinesterase inhibition
data for risk assessment purposes. (U.S. EPA, 1993)
Multiple adverse effects resulting from cholinesterase inhibition and other toxic
mechanisms have been associated with organophosphate exposure. The SAB
report states that:
Clinical effects associated with exposure to cholinesterase inhibitors can
be used in risk assessment to define hazard and to calculate benchmark
doses and RfDs. (U.S. EPA, 1993)
Although the issue may be clearer for clinical effects, there has not, as yet, been
resolution of the question as to whether cholinesterase inhibition alone should be
used as a critical endpoint. The 1993 report "does not provide a simple yes or no
answer to the issue of using RBC cholinesterase inhibition data by itself (i.e., in the
absence of clinical symptoms) for risk assessment." Concern arises, in part, from
the fact that blood enzyme inhibition may precede and predict brain enzyme
inhibition, which is of significant concern (U.S. EPA, 1993). Additional information
is required to clarify this issue.
C-5
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APPENDIX C
The SAB report suggests that EPA continue research designed to.evaluate the
correlation of clinical signs with blood cholinesterase inhibition, especially
correlations with respect to dose, time, and linearity (U.S. EPA, 1993).
Cholinesterase inhibition has been used as the critical endpoint, on which RfD
calculations are based, for many organophosphates included in the IRIS database.
Because of the uncertainty surrounding the use of cholinesterase inhibition as an
effect, readers may wish to calculate their own exposure limits. When the RfD for
the target analytes is based on cholinesterase inhibition, other chronic toxicity data
are provided for analytes having sufficient data. This enables the reader to
calculate estimated exposure limits (using Equations 3-1 and 3-3 in Section 3 of
this report) and derive fish consumption limits (using Equation 3-2 in Section 3)
based on other health endpoints if appropriate.
Effects commonly associated with organophosphate exposure 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, 1993; U.S. EPA, 1982). In addition,
some, but not all, organophosphates cause peripheral neuropathy resulting from
demyelination of the nerves. Specific effects include numbness, tingling, pain,
weakness, and paralysis in the arms and legs. These effects may be delayed and
may be reversible or irreversible (U.S. EPA, 1982).
Muscarinic effects in children exposed to organophosphates may differ from those
in adults. Those most commonly encountered with acute exposure include: CNS
depression, stupor, flaccidity, dyspnea, and coma. Seizures may be more common
in children than in adults (HSDB, 1993).
Psychiatric symptoms that have been reported include defects in expressive
language and cognitive function, impaired memory, depression, anxiety or
irritability, and psychosis. These are more common in individuals with other clinical
signs of organophosphate poisoning or with preexisting psychological conditions
(HSDB, 1993). Behavioral effects are a prominent concern based on the results
of toxicity data reviewed for the target analytes. It appears to be one of the most
sensitive indicators of toxicity related to chronic exposure. Behavioral effects, such
as aggressiveness, irritability, and hyperactivity, occurred at low levels of exposure
in animal studies. These effects are particularly problematic because they are
difficult to specifically associate with organophosphate exposure in the human
population but could have serious consequences.
There is a recognized high-risk human population with respect to organophos-
phate 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
C-6
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APPENDIX C
... . cholinesterase activities. Red blood cell (RBC) acetylcholinesterase is reduced in
certain conditions such as hemolytic anemias; people with these conditions are at
greater risk than the general population from exposure to organophosphates (U.S.
EPA, 1982).
Compounds known to reduce plasma pseudocholinesterase activity and thereby
aggravate the effects of cholinesterase inhibitors are carbon disulfide,
benzalkonium salts, organic mercury compounds, ciguatoxins, and solanines (U.S.
EPA, 1982).
The Hazardous Substances Data Bank (HSDB, 1993) contains a summary of
diseases and disorders that are of special concern for individuals exposed to
organophosphates. The following are listed as contraindications for work with (and
exposure to) these chemicals: "organic diseases of the CNS, mental disorders and
epilepsy, pronounced endocrine and vegetative disorders, pulmonary tuberculosis,
bronchial asthma, chronic respiratory diseases, cardiovascular diseases and
circulatory disorders, gastrointestinal diseases (peptic ulcer), gastroenterocolitis,
diseases of the liver and kidneys, and eye diseases (chronic conjunctivitis and
keratitis)" (HSDB, 1993).
C.3 REFERENCES
ATSDR (Agency for Toxic Substances and Disease Registry). 1992a.
Toxicological Profile for Chlordane. U.S. DHHS, PHS, Atlanta, GA.
_. 1992b. Draft Toxicological Profile for ODD, DDT, DDE. U.S. DHHS,
PHS, Atlanta, GA.
Cunningham-Burns, K.M., and W.H. Hallenbeck. 1986. Pesticides and Human
Health. New York: Springer-Verlag Press.
Hayes, W.J. 1982. Pesticides Studied in Man. Baltimore, MD: Williams and
Wilkins.
Hayes, W.J., and E.R. Laws. 1991. Handbook of Pesticide Toxicology, Vols. 1-3.
San Diego: Academic Press, Inc.
HSDB (Hazardous Substances Data Bank). 1993. All searches conducted on-line
through Toxnet in 1993 unless specifically noted. Developed by U.S.
Environmental Protection Agency.
Klaassen, C.D., M.O. Amdur, and J. Doull (eds.). 1986. Casarett and Doull's
Toxicology: The Basic Science of Poisons. New York: Macmillan Publishing.
U.S. EPA (Environmental Protection Agency). 1982. Recognition and Manage-
ment of Pesicide Poisonings, 3rd Ed. EPA-540/9-80-005. Washington, DC:
U.S. Government Printing Office. January.
C-7
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APPENDIX C
U.S. EPA (Environmental Protection Agency). 1993. A SAB Report:
Cholinesterase Inhibition and Risk Assessment, Review of the Risk Assess-
ment Forum's Draft Guidance on the Use of Data on Cholinesterase Inhibition
in Risk Assessment by the SAB/SAP Joint Committee. Washington, DC.
C-8
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APPENDIX D
POPULATION EXPOSURE ASSESSMENT-
CONSUMPTION PATTERNS AND SURVEYS
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APPENDIX D
APPENDIX D
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.
D.1 HIERARCHY OF FISH CONSUMPTION INFORMATION
Table D-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
of fish 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
number of the survey respondents were Asian-American. If the survey was
D-3
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APPENDIX D
Table D-1. Hierarchy of Data Sources5
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'3
• 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)
• U.S. Department of Agriculture Continuing Survey of Food Intake by Individuals
(CSFII) studies
• Other national surveys that estimate fish consumption patterns
• Census data
a 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 or State
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).
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.
D.1.1 Local Fish Consumption Data
D.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
agencies involved with fish and wildlife management perform creel surveys or
censuses. These surveys consist of clerks interviewing fishers onsite and
D-4
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APPENDIX D
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.
D.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.
D.1.2 Regional or State Consumption Data
D.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 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
D-5
-------
APPENDIX D
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.
D.1.2.2 Behavioral Risk Surveillance Surveys-
Most States already participate in random telephone surveys under the Behavioral
Risk Surveillance System (BRSS). The BBSS 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).
D.1.3 National Consumption Data
D.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 (U.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.
D.1.3.2 U.S. Department of Agriculture CSCFIl Study—
The Continuing Survey of Food Intake by Individuals (CSFII) is a national food
consumption survey conducted annually by the USDA. In the CSFIIs, dietary
intake data collection is distributed over a year long period from a sample of
individuals in the 48 conterminous States (USDA, 1991). Survey participants
provide 3 consecutive days of data. On the first day of the survey, participants
D-6
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APPENDIX O
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.
D.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, Consumption
Surveys for Fish and Shellfish: A Review and Analysis of Survey Methods (U.S.
EPA, 1992), 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.
D.2.1 Key Considerations
Table D-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
Table D-2. Key Considerations for Effective
Fish Consumption Surveys
Population Selection
Population Access
Consumption Rates
Consumption Patterns
Duration of Study
What population is to be surveyed?
Based on what criteria (e.g., jurisdictional region, region with
known fish contamination)?
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)?
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)?
How are variations in consumption patterns accommodated
(e.g., preparation methods, type of fish eaten, parts of fish
consumed)?
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?
D-7
-------
APPENDIX D
credibility of the survey results must be ensured through careful survey prepara-
tion, 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
• 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 subpopu-
lations at high risk, such as urban low-income populations or individuals of a par-
ticular 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.
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
D-8
-------
APPENDIX D
either "how many meals of fish (or what amount of fish) have you (and household
members) eaten in this past week?" or "how many meals of fish (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).
D.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
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.
D-9
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APPENDIX D
• 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 fish 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.
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.
D-10
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APPENDfX D
D.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 2.4.3.1, 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,
liyer, 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.
D.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
D-11
-------
APPENDIX D
mg/g of fish tissue, and 1,000 g of fish 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)
(D-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 3-3 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
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) .
(D-3)
Information regarding the duration and periodicity of exposure is needed for both
determining potential risks and identifying the most appropriate consumption limits.
D-12
-------
APPENDIX D
It should be described when exposure information is presented for use in risk
assessment.
D.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 D-3 through D-6. They are presented by Region,
proceeding from east to west across the United States.
Tables D-3 and D-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 are included with the consumption
rate, because they directly impact 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.
Survey methods used to collect the data reported in Tables D-3 and D-5 are listed
in Tables D-4 and D-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
D-13
-------
APPENDIX D
Table D-3. Sport Fishers8 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 (1 0
counties)8
Ontario fishers9
Los Anaeles Harbor
Mean
45.8
28.1
40.9
14.5
18.3
44.7
12.3
26.1
22.5
Consumption Rates (g/d)
80th 90th 95th
Median Percentile Percentile Percentile Fish Type
50.7 F+S, F+C
65 F+S, F+C
F+S, R+C
F+S, R
30 62 80 F+S, R
»50 F+S, R+C
F, R
37.3 F, R
63.4 F, R+C
F,R
37 225 S, R
fishers10
Washington State
(Commencement Bay)
fishers11
Washington State
(Columbia River) fishers
Maine fishers (inland
waters)13
23
54
12
7.7
6.4
2.0
13
26
SOURCES:
1 ALDEM(1993).
2 Dellenbargeretal. (1993).
3 Connelly etal. (1990).
4 Barclay (1993).
5 West etal. (1993).
6 West etal. (1989).
7 Humphrey (1976).
8 Fioreetal. (1989).
9 Cox etal. (1993).
10 Puffer etal. (1982).
11 Pierce etal. (1981).
12 Honstead etal. (1971).
13 Ebertetal. (1993).
S,R
F+S, R+C
F,R
F = freshwater, S = saltwater, R = recreationally caught, C = commercially caught.
* Sport fishers may include individuals who eat sport-caught fish as a large portion of their diets.
D-14
-------
APPENDIX D
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-------
APPENDIX D
Table D-5. Subsistence Fishers3 Consumption Data
Consumption Rates (g/d)
Fisher Group
Great Lakes Tribes1
Columbia River Tribes2
High-end Caucasian consumers on
Lake Michigan3
Mean
351
58.7
48b
27°
109
95th percentile Max
1,426
170
144
132
Fish Type
F
F
F
F
F+S
F = fish, S = shellfish.
* 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.
0 Data from 1989 survey of fish eaters.
SOURCES:
1 Kmiecik(1994).
2 CRITFC(1994).
3 Hovinga (1992,1993).
4 Nobmannetal. (1992).
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.
D.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 D-3
and the survey methods used to collect the data are summarized in Table D-4.
D-16
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D-17
-------
APPENDIX D
D.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.
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
D-18
-------
APPENDIX O
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 D-5 and the survey methods
used to collect the data are summarized in Table D-6.
D.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 D-7. The
details of the survey methods used in these studies are summarized in Table D-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.
D.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
D-19
-------
APPENDIX D
Table D-7. National Studies Consumption Data
Consumption Rates (g/d)
Population
Mean
95th Percentile
Fish Type
US1
us2
us2
us3
6.6 47.3
6.5 —
14.3 41.7
16.7
F+E, C+R
F+E, C+R
F+S, C+R
F+S, 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, 1980).
3 Market Facts Survey (as cited in Javitz, 1980).
general population (assuming a consumption rate of 6.5 g/d). Many of the
contaminants examined in Section 5 of this volume have developmental 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
• • available. In the absence of these data, U.S. census-data on average family size
canbeused. '
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.
D-20
-------
APPENDIX D
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D-21
-------
APPENDIX D
D.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 E) 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).
D.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., March.
Carlson, G. 1994. Comments on Volume 2, Risk Assessment and Fish Consump-
tion Limits (first edition) from the Missouri Department of Health. April 22.
-------
APPENDIX D
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. Ontario, Canada: Ministry of Environment and
Energy. November.
CRITFC (Columbia River Inter-Tribal Fish Consumers). 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.
Cung, Josee. 1994. Minnesota Department of Natural Resources. South East
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.
Dourspn, M.L., and J.M. Clark. 1990. Fish consumption advisories: Toward a
unified, scientifically-credible approach. Regulatory Toxicity and Pharma-
cology12.
Ebert, E.S., N.W. Harrington, K.J. Boyle, J.W. Knight, R.E. Keenan.
Estimating consumption of freshwater fish among Maine anglers.
American Journal of Fisheries Management. 13(4):737-745.
1993.
North
Fiore, B.J., et al. 1989. Sport fish consumption and body burden levels of
chlorinated hydrocarbons: A study of Wisconsin anglers. Arch Env Health
44:82-88.
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. [Cited in Rupp
etal., 1980.]
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. Archives
of Environmental Contamination and Toxicology 22(4):362-366. May.
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. Archives of Environmental Health 48(2):98-104. May
D-23
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APPENDIX D
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.
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
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, and S.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. U.S. Environmental Protection Agency, Environmental Research
Laboratory, Corvallis, OR. EPA 600/3-82-070.
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.
Rupp Elizabeth, F.L. Miller, and I.C.F. Baes III. 1980. Some results of recent
surveys of fish 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. Toxicology and Industrial
Health 12 (3-4):427-434.
D-24
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APPENDIX D
University of Wisconsin SeaGrant. 1994. Personal communication with Abt
Associates. May 27.
USDA (U.S. Department of Agriculture). 1991. Continuing Survey of Food Intakes
by Individuals Data and Documentation. Human Nutrition Information Service
Belcrest Road, Hyattsville, MD.
U.S. EPA (Environmental Protection Agency). 1988. Region V Risk Assessment
for Dioxin Contaminants. Chicago, IL.
1992. Consumption Surveys for Fish and Shellfish: A Review and
Analysis of Survey Methods. Washington, DC: Office of Water.
_. 1995. Guidance for Assessing Chemical Contaminant Data for Use in
Fish Advisories. Volume 1: Fish Sampling and Analysis, Second Edition.
Washington, DC: Office of Science and Technology, Office of Water.
U.S. FWS (Fish and Wildlife Service). 1993. National Survey of Fishing, Hunting,
and Wildlife Associated Recreation. Washington, DC.
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.
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.
Young, Pat. 1994. U.S. Environmental Protection Agency Region 9. Personal
communication with Abt Associates, July 28.
D-25
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-------
APPENDIX E
DOSE MODIFICATIONS DUE TO FOOD PREPARATION AND COOKING
-------
-------
APPENDIX E
APPENDIX E
DOSE MODIFICATIONS DUE TO FOOD PREPARATION AND COOKING
E.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 to account for preparation and cooking
into the exposure equation requires two types of information:
• Methods used by fish consumers to prepare (trim, skin) and cook (broil, bake,
barbeque, fry, smoke) 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 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 modification
factor for removing skin if the fish concentrations 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
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.
E.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 differing exposure levels to chemicals. This section is meant as an overview;
States should consult primary research studies for more information. In general,
contaminant concentrations differ among
• Fatty tissues, muscle tissue, and internal organs
• Different species of fish
E-3
-------
APPENDIX E
Different age or size classes of fish
Type of chemical contaminant present in the fish.
E.1.2 Fish Tissue Types
Lipophiiic chemicals accumulate mainly in fatty tissues, including the belly, 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 concen-
trations than fatty tissues (Great Lakes Sport Fish Advisory Task Force, 1993) but
contains more mercury, which binds to proteins (Minnesota Department of Health,
1992).
Many people remove the internal organs before cooking fish and trim off fat 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 (Dellinger, 1996). Certain subpopula-
tions, including some Asian 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. 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.
E.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 bio-magnifica-
tion, 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 waterbody also determines its relative bioaccumulation
potential. Bottom feeders, such as carp or catfish, are exposed to more sediments
than are fish that feed in 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 toxicants than fattier fish such as bluefish, rainbow trout, lake
E-4
-------
APPENDIX E
trout, some salmon, catfish, and carp. Aquatic organisms also differ in their abilities
to metabolize and excrete contaminants. For example, one study found fish more
able to metabolize benzo[a]pyrene than shrimp, amphipod crustaceans, and
clams, respectively (U.S. EPA, 1995a). The ability to break down and excrete
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 be more highly exposed to organics than an
individual who eats primarily leaner fish species. Thus, States should consider
multiple species exposure in their decision to issue fish consumption advisories.
E.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 (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 in public
education efforts.
E.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 (Stern et al., 1 992). Thus, fish exposed to similar concentrations of
different contaminants may accumulate differing levels of contaminants in their
tissues. .
E.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 tissues compared with
the concentration in the whole untrimmed fish. Cadmium concentrates largely in
the liver, followed by the kidneys and gills, and less so in the muscle tissue
(Harrison and Klaverkamp, 1990; Jaffar and Ashraf, 1988; Marcovecchio et al.,
1988; Norey et al., 1990), indicating that cadmium concentrations could be
E-5
-------
APPENDIX E
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). Although all three heavy metals are bioaccumulative, cadmium
and mercury were found to bioaccumulate at higher rates in some species than in
others (Jaffarand Ashraf, 1988).
E.1.5.2 Organochlorines—
Organochlorine pesticides and PCBs tend to concentrate in fatty tissues (Branson
et al., 1985; Kleeman et al., 1986a, 1986b; Ryan et al., 1983; U.S. EPA, 1995a).
One study positively correlated PCB and mirex levels with fat levels across 10
freshwater fish species (Ryan et al., 1983). These compounds are neither readily
metabolized nor excreted and thus tend to bioaccumulate through the food web'
(U.S. EPA, 1995a). As fish species store fat differently, so will they concentrate
organochlorines differently.
PCB levels have been studied in several species and tissues of fish. Adult rainbow
trout were found to store PCBs in the carcass and in skeletal muscle, while adult
and juvenile yellow perch stored PCBs in the viscera and carcass (Kleeman et al.,
1986b). Higher chlorinated biphenyls have been found to bioaccumulate more
readily than lower chlorinated biphenyls (Bruggeman et al., 1984; U.S. EPA,
1995a). Unfortunately, some of these higher chlorinated biphenyls tend to be the
more potent toxicants as well (Williams et al., 1992).
E.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 no
information was found as to how they accumulate differentially in fish tissues.
Organophosphates as a group have similar chemical characteristics although they
are less persistent in the environment than the organochlorines (U.S. EPA, 1995a).
States may wish to use this chemical-specific information on distribution 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 of the fish, or eating only fat or lean fish species.
E.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 for analyzing chemical
contaminants. Readers are referred to Volume 1, 2nd edition, of this series for a
more complete discussion of this analysis (U.S. EPA, 1995a). The fillet should
consist of the portion of the individual organism commonly consumed by the
E-6
-------
APPENDIX E
general U.S. population or a specific subpopulation of concern. 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
subpopulation typically consumes whole fish. Following these guidelines, States
may have concentration data from samples with skin-on or from whole fish. In food
preparation, fish may be further trimmed and have additional fat removed.
When States have data on the preparation methods of the target fish consuming
populations, appropriate 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 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 of fish, including Native American and
Asian American fishers, often consume most of the fish and may do very little
trimming. Consequently, assuming a loss of toxic chemicals may lead to an
underestimate of exposure and risk for these groups.
EPA does not recommend the use of dose modification factors for
setting health-based intake limits unless data on local methods of prep-
aration and their impact on contaminant concentrations are available.
EPA does, however, recommend that all fish advisories emphasize the importance
of skinning, trimming (including gutting), and certain ways of cooking as effective
means to minimize the risks from chemical contaminants. 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 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 the
• fish species and size (age class)
• chemical contaminant
• specific food preparation and cooking techniques used.
The results of a number of fish preparation and cooking studies are presented in
Tables E-1 and E-2. The data are relevant primarily to concentrations in the
standard fillet. Dose modification will depend on how the dose is determined
E-7
-------
APPENDIX E
Table E-1. Summary of Contaminant Reductions Due to Skinning, Trimming, and
Cooking (Based on Standard Fillet)
Species Contaminant
Brown Trout DDE
DDE
DDE
Mirex
Mirex
Mirex
Mirex
Mirex
PCB
PCB
PCB
PCB
PCB
Carp cc-Chlordane
a-Chlordane
a-Chlordane
a-Chlordane
Dieldrin
Dieldrin
Dieldrin
Dieldrin
Heptachlor epoxide
PCB
PCB
PCB
PCB
Chinook a-Chlordane
Salmon a-Chlordane
a-Chlordane
a-Chlordane
a-Chlordane
a-Chlordane
a-Chlordane
Dieldrin
Dieldrin
Dieldrin
Dieldrin
Dieldrin
Dieldrin
Dieldrin
Activity3 Reduction (%)b
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 & 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
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
31
44
41
45
37
27
42
51
30
31
40
40
29
40
50
Reference
Skeaetal. (1979)
Skeaetal. (1979)
Skeaetal. (1979)
Voilandetal. (1991)
Skeaetal. (1979)
Skeaetal. (1979)
Skeaetal. (1979)
Skeaetal. (1979)
Voilandetal. (1991)
Skeaetal. (1979)
Skeaetal. (1979)
Skeaetal. (1979)
Skeaetal. (1979)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
See footnotes at end of table.
(continued)
E-8
-------
APPENDIX E
Table E-1. (continued)
Species Contaminant
Chinook Heptachlor epoxide
Salmon (con.) 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
Lake Trout a-Chlordane
a-Chlordane
a-Chlordane
a-Chlordane
DDT
DDT
DDT
DDT
Dieldrin
Dieldrin
Dieldrin
Dieldrin
Heptachlor epoxide
Heptachlor epoxide
Heptachlor epoxide
Heptachlor epoxide
PCB
PCB
PCB
See footnotes at end of table.
Activity3 Reduction (%)b Reference
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 & 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
Skin-off & salt boiling
Skin-on & smoking
Skin-off & baking
Skin-off & charbroiling
Skin-off & salt boiling
52
40
42
37
23
45
48
38
44
46
36
33
40
49
34
30
34
74
22
37
47
26
41
6
53
14
21
1
60
8
15
16
43
39
39
3
59
13
29
10
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabik et al. (1 993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
(continued)
E-9
-------
APPENDIX E
Table E-1. (continued)
Species Contaminant
Lake Trout PCB
(con.) Toxaphene
Toxaphene
Toxaphene
Toxaphene
Smallmouth DDE
Bass DDE
DDE
Mirex
Mirex
Mirex
Mirex
PCB
PCB
PCB
PCB
Walleye DDT
DDT
DDT
cc-Chlordane
cc-Chlordane
a-Chlordane
Dieldrin
Dieldrin •
Dieldrin
PCB
PCB
PCB
Toxaphene
Toxaphene
Activity3 Reduction (%)b
Skin-on & smoking
Skin-off & baking
Skin-off & charbroiling
Skin-off & salt boiling
Skin-on & smoking
Trimming
Baking
Frying
Trimming
Baking
Frying
Trimming & cooking
Trimming
Baking
Frying
Trimming & cooking
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 & charbroiling
46
31
40
5
51
54
16
75
64
21
75
80
64
16
74
80
4
16
11
32
33
-25
3
3
26
17
24
14
45
43
Reference
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Skeaetal. (1979)
Skeaetal. (1979)
Skeaetal. (1979)
Skeaetal. (1979)
Skeaetal. (1979)
Skeaetal. (1979)
Skeaetal. (1979)
Skeaetal. (1979)
Skeaetal. (1979)
Skeaetal. (1979)
Skeaetal. (1979)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
Zabiketal. (1993)
• 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.
b Data from the Zabik (1993) study were condensed by averaging contaminant reductions across lakes
whenever a fish species was sampled from more than one of the Great Lakes.
E-10
-------
APPENDIX E
Table E-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
••- . PGB--- - "
" PCB . . • :
Mirex
PCB
PCB (1248)
PCB (1248)
PCB (1254)
PCB (1254)
DDT
DDT/DDE
DDT
Mirex
PCB
PCB (1248)
PCB (1248)
PCB (1254)
PCB(1254)
Dieldrin
Dieldrin
DDT
DDT/DDE
DDT
DDT
DDT
DDT
DDT
DDT
Dieldrin
Mirex
PCB
DDT
PCB
PCB
PCB
Activity
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
Reduction (%)a
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
Reference
NYSDEC(1981)
NYSDEC(1981)
Armbruster et al. (1989)°
Armbruster et al. (1989)°
Armbruster et al. (1989)°
Armbruster efal. (1989)°
Armbruster et'al. (1989)5
Armbruster et al. (1989)°
NYSDEC(1981)
NYSDEC(1981)
Smith etal. (1973)
Smith etal. (1973)
Smith etal. (1973)
Smith etal. (1973)
Reinert etal. (1972)
NYSDEC(1981)
Reinert etal. (1972)
NYSDEC(1981)
NYSDEC(1981)
Smith etal. (1973)
Smith etal. (1973)
Smith etal. (1973)
Smith etal. (1973)
Zabik etal. (1993)
Zabik etal. (1993)
Reinert etal. (1972)
NYSDEC(1981)
Reinert etal. (1972)
Reinert etal. (1972)
Reinert etal. (1972)
Zabik etal. (1993)
Zabik etal. (1993)
Zabik etal. (1993)
Zabik etal. (1993)
NYSDEC(1981)
NYSDEC(1981)
Reinert etal. (1972)
EPA (1992)
EPA (1992)
EPA (1992)
a It could not be positively determined that reduction figures were calculated as changes in contaminant
concentrations from the standard fillet.
b Average of findings reported in New York State Department of Environmental Conservation (1981) and White et
al. (1985).
0 Averages of findings reported in Armbruster et al. (1989).
E-11
-------
APPENDIX E
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 methods.
Table E-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 E-2. Both tables 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 (Voiland
et al., 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 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 iiquefication 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 E-1 through E-3 do not address chemical
degradation due to heat applied in cooking. Zabik et al. (1993) found that smoking
Table E-3. Average Contaminant Reductions Due to Cooking in Great
Lakes Fish a
Chemical Contaminant
Reduction (%)
p.p'-DDT
p.p'-DDE
p,p'-DDD
a-Chlordane
v-Chlordane
Oxychlordane
c/s-Nonachlor
frans-Nonachlor
Dieldrin
Heptachlor epoxide
Toxaphene
Total PCBs
34.0
29.4
29.0
34.8
33.0
35.6
35.7
27.9
28.7
35.6
36.5
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. (1993).
E-12
-------
APPENDIX E
lake trout reduced pesticides and total RGBs significantly more than other cooking
methods, but resulted in the formation of low levels of PAHs. Until there is more
information about the toxicity of the byproducts generated during the degradation
of PCBs or the other chemicals of concern, EPA recommends that no dose
modification be assumed due to degradation alone.
Zabik et al. (1993) 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 E-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 E-3. The information provided in this table is not
species-specific, which may limit the situations to which it is applicable.
E.3 REFERENCES
Armbruster, G., K.L. Gall, W.H. Gutenmann, and D.J. Lisk. 1989. Effects of
trimming and cooking by several methods on polychlorinated biphenyls
(PCBs) residues in bluefish. J. Food Safety 9:235-244.
Branson, Dean R., IT. Takahashi, W.M. Parker, and G.E. Blau. 1985.
Bioconcentration kinetics of 2,3,7,8-tetrachlorodibenzo-p-dioxin in rainbow
trout. Environmental Toxicology and Chemistry 4.
Bruggeman, W.A., A. Opperhuizen, A. Wijbenga, and O. Hutzinger. 1984.
Bioaccumulation of superlipophilic chemicals in fish. Toxicology and
Environmental Chemistry 7.
Dellinger, J.A. 1996. Department of Preventative Medicine, Medical College of
Wisconsin. Personal communication with Abt Associates. March 26.
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.
Harrison, S.E., and J.F. Klaverkamp. 1990. Metal contamination in liver and
muscle of northern pike (Esox lucius) and white sucker (Catostomus
commersoni) and in sediments from lakes near the smelter at Flin Flon,
Manitoba. Environmental Toxicology and Chemistry 9.
Jaffar, M., and M. Ashraf. 1988. Heavy metal contents in some selected local
freshwater fish and relevant waters. Indian Journal of Marine Science 17(3).
Kleeman, James, J.R. Olson, S.S. Chen, and R.E. Peterson. 1986a. Metabolism
and disposition of 2,3,7,8-tetrachIorodibenzo-p-dioxin in rainbow trout.
Toxicology and Applied Pharmacology 83.
E-13
-------
APPENDIX E
Kleeman, James, J.R. Olson, S.'.S. Chen, and R.E. Peterson. 1986b. 2,3,7,8-
tetrachlorodibenzo-p-dioxin metabolism and disposition in yellow perch.
Toxicology and Applied Pharmacology 83.
Marcovecchio, J.E., V.J. Moreno, and A. Perez. 1988. The sole, paralichthys sp.,
as an indicator species for heavy metal pollution in the Bahia Blanca Estuary,
Argentina. Science of the Total Environment 75.
Minnesota Department of Health. 1992. Minnesota Fish Consumption Advisory.
Minneapolis, MN. May.
NYSDEC (New York State Department of Environmental Conservation). 1981.
Toxic Substances in Fish and Wildlife. Technical Report 81-1 (BEP). Albany,
NY: Division of Fish and Wildlife.
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. Comparative Biochemical Physiology C., Vol.
96C,No.1,
Reinert, R., et al., 1972. Effects of dressing and cooking on DDT concentrations
in certain fish from Lake Michigan. J Fish Res Board Can 29.
Ryan, J., P. Lau, J. Pilon, and D. Lewis. 1983.2,3,7,8-Tetrachlorodibenzo-p-Dioxin
and 2,3,7,8-Tetrochlorodibenzofuran Residues in Great Lakes Commercial
and Sport Fish. In G. Gehoudhar, L. Keithand, C. Rappe (eds.). Chlorinated
Dioxins and Dibenzofurans in the Total Environment. Boston, MA: Butterworth
Pub.
Skea, J.C., et al. 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).
Smith, W.E., K. Funk, and M.E. Zabik. 1973. Cited in: Assessment of
Contaminants in Five Species of Great Lakes Fish at the Dinner Table. Final
Report to the Great Lakes Protection Fund. March.
Stern, G., G. Muir, C. Ford, N. Grift, E. Dewally, T. Bidleman, and M. Walls. 1992.
Environ Sci Technol. 26:1838-1840.
U.S. EPA (Environmental Protection Agency). 1992. National Study of Chemical
Residues in Fish, Volumes I and II. EPA 823-R-92-008a. Washington, DC:
EPA, Office of Science and Technology.
. 1995a. Guidance for Assessing Chemical Contamination Data for Use
in Fish Advisories, Volume 1: Fish Sampling and Analysis. Second Edition.
Washington, DC: Office of Science and Technology.
E-14
-------
APPENDIX Ł
U.S. EPA (Environmental Protection Agency). 1995b. Guidance for Assessing
Chemical Contamination Data for Use in Fish Advisories, Volume 4: Risk
Communication. Washington, DC: Office of Science and Technology EPA-
823-R-95-001.
Voiland Jr., M.P., et al. 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).
White, R., et al. 1985. PCBs in striped bass collected from the Hudson River, NY,
during fall 1981. JEnvirContam Tox/co/34.
Williams, Lisa, J.P. Glesy, N. DeGalan, D.A. Verbrugge, D.E. Tillitt, G.T. Ankley,
and R.L. Welch. 1992. Prediction of Concentrations of 2,3,7,8-tetrachloro-
dibenzo-p-dioxin equivalents from total concentrations of polychlorinated
biphenyls in fish fillets. Environmental Science and Technology 26(6).
Zabik, M.E., et al. 1993. Assessment of Contaminants in Five Species of Great
Lakes Fish at the Dinner Table. Final Report to the Great Lakes Protection
Fund, March.
E-15
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APPENDIX F
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.
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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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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.
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
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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.
Kev 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
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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.Effe.C-t
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.
APPROVED:
DATE:
MAR 2 11995
Carol M. Browrv
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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 standalone; 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 carcinogenicity; 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.
B.
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?
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?
. 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
n. Risk Assessment and Risk Characterization
ffl. 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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H. 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 uncertainty); 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 in 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. Pose-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 TTT 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 qj* 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.
In 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
High 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 interpcetable 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
'^Maximizing 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
10'6? . .. .
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
However, 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).
Li 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.
*7
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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»U.S. GOVERNMENT PRINTING OFFICE: 1997-521-099/90227
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