Office of the Science Advisor

STAFF PAPER
EPA/100/B-04/001 

March 2004 

U.S. Environmental Protection Agency


AN EXAMINATION OF EPA RISK


ASSESSMENT PRINCIPLES AND


PRACTICES


Staff Paper Prepared for the U.S. Environmental Protection Agency 

by members of the Risk Assessment Task Force 

Office of the Science Advisor 

U.S. Environmental Protection Agency 

Washington, DC 20460 
Page ii Examination of EPA Risk Assessment Principles and Practices 

DISCLAIMERS 

This document has been reviewed in accordance with United States Environmental 

Protection Agency policy and approved for publication and distribution. Mention of trade names 

or commercial products does not constitute endorsement or recommendation for use. 

This document presents an analysis of EPA’s general risk assessment practices, based on 

typical historic and current practice. The document does not establish new Agency policy or 

guidance or amend any existing Agency policy or guidance. Nor does the document attempt to 

present binding prospective requirements, necessarily applicable to future agency actions. The 

use of the words “should,” “can,” “would,” and “may” in this document means that something is 

suggested or recommended, but not required. 

A particular risk assessment practice described in this document may not apply to an 

individual situation based upon the circumstances. Interested parties are free to raise questions 

and objections about the substance of the practices discussed in this document and the propriety 

of the application of those practices to a particular situation. Any individual or site-specific risk 

management decision will be based on the applicable statute and regulations, and on facts 

specific to the circumstances at issue. Variance from the approaches outlined in this document 

does not necessarily have any significance. EPA and other decision makers retain the discretion 

to adopt approaches on a case-by-case basis that differ from those described in this document 

where appropriate. 

Risk assessments discussed in this staff paper reflect a “snapshot” in time and may not be 

reflective of any further assessment activity past the time of a particular description. For 

example, the Integrated Risk Information System (IRIS) descriptions, particularly of past 

assessments, may not be reflective of the current IRIS data base, as assessments are continuously 

updated. Further, it is important to note that current IRIS health assessments are conducted using 

the 1999 draft cancer guidelines (as of this examination), and not the 2003 draft final cancer 

guidelines. 
Examination of EPA Risk Assessment Principles and Practices Page iii

TABLE OF CONTENTS


ACKNOWLEDGMENTS .................................................. Page ix

AN EXAMINATION OF EPA RISK ASSESSMENT PRINCIPLES


 AND PRACTICES ................................................. Page 1

1. INTRODUCTION TO EPA RISK ASSESSMENT ............................ Page 2

1.1	 Overview ................................................... Page 2

1.1.1	 What Is Risk Assessment? ................................ Page 2

1.1.2	 Why Does EPA Conduct Risk Assessments? ................. Page 2

1.1.3	 How Does EPA Use Risk Assessments in Decision Making? .... Page 3

1.1.4	 What Is Some Historical Perspective Relevant to EPA


Risk Assessment Practices? ............................... Page 4

1.2	 EPA Process for Evaluation of Risk Assessment Principles and Practices . Page 6

1.2.1	 Why Are We Conducting an Evaluation of Our Risk Assessment


Principles and Practices Now? ............................. Page 6

1.2.2	 What Is the Purpose and Intent of This Staff Paper? ............ Page 6

1.2.3	 Through What Process Is EPA Conducting This Evaluation? ..... Page 7

1.2.4	 How Does the 2003 OMB Draft Report to Congress Relate to Our


Evaluation? ........................................... Page 7

1.2.5	 What Is the General Nature of the Comments on EPA Risk


Assessment Practices Submitted to OMB? ................... Page 8

1.3	 Organization of This Document .................................. Page 8

1.3.1	 How Are the General Comments Addressed? ................. Page 8

1.3.2	 How Are the Specific Comments and Examples Addressed? ..... Page 9

1.4	 Other Components Impacting Risk Assessment, But Not Addressed


in This Document ............................................. Page 9

1.4.1	 What Impact Do the OMB and EPA Information Quality


Guidelines Have on EPA Risk Assessment Practices? .......... Page 9

1.4.2	 Why Is Peer Review So Important? ........................ Page 10

2. EPA RISK ASSESSMENT AND PUBLIC AND ENVIRONMENTAL HEALTH


PROTECTION .................................................... Page 11

2.1	 Public and Environmental Health Protection (“Public Health


Protection”) ................................................ Page 11

2.1.1	 What Is EPA’s General Approach for Developing


Risk Assessments? ..................................... Page 11

2.1.2	 What Is the “Conservatism” Issue in Terms of Public


Health Protection? ..................................... Page 11
Page iv Examination of EPA Risk Assessment Principles and Practices 

2.1.3	 Is Science Policy Utilized Within the Risk Assessment


Process? ............................................. Page 12

2.1.4	 What Impact Does Statutory Language Have on EPA Risk


Assessment Practices for Public Health Protection? ........... Page 13

2.1.5	 Does EPA Take a Reasonable Approach to Public Health


Protection? ........................................... Page 16

2.1.6	 What Happens When Default Assumptions Are Combined? .... Page 17

2.2	 General Risk Assessment Approaches Used by EPA for Public


and Environmental Health Protection (“Public Health Protection”) ..... Page 20

2.2.1	 How Comprehensive Are EPA’s Risk Assessments? .......... Page 20

2.2.2	 Whom Is EPA Trying To Protect? ......................... Page 22

2.2.3	 Are Risk Assessment and Risk Management Separate? ........ Page 22

2.2.4	 How Do “Planning and Scoping” Help Environmental


Risk Assessment? ..................................... Page 23

2.2.5	 How Does EPA Use a Screening Risk Assessment? ........... Page 24

2.2.6	 What Happens if EPA Identifies a Potential Risk That Needs


To Be Addressed After a Screening Risk Assessment? ......... Page 25

2.2.7	 How Are High-End Exposures Reflected in EPA Evaluations? .. Page 26

3. UNCERTAINTY AND VARIABILITY .................................... Page 30

3.1	 Overview .................................................. Page 30

3.2	 Uncertainty and Variability .................................... Page 30

3.2.1	 What Is Uncertainty? ................................... Page 30

3.2.2	 What Is Parameter Uncertainty? .......................... Page 31

3.2.3	 What Is Model Uncertainty? ............................. Page 32

3.2.4	 How Does Variability Differ From Uncertainty? ............. Page 32

3.3	 Characterizing Uncertainty and Variability ........................ Page 33

3.3.1	 Why Is Characterizing Uncertainty and Variability Important? .. Page 33

3.3.2	 When Should EPA Conduct Uncertainty Analysis? ........... Page 34

3.3.3	 What Is an Appropriate Level of Uncertainty Analysis? ........ Page 35

3.3.4	 With What Precision Should EPA Results Be Reported? ....... Page 36

3.4	 Issues in Characterizing Uncertainty and Variability ................. Page 36

3.4.1	 What Is the Issue of Deterministic Versus Probabilistic


Approaches? .......................................... Page 37

3.4.2	 What Is the Importance of Quantitative Characterization


of Uncertainty in Dose-Response? ......................... Page 38

3.4.3	 How Does EPA Use Probabilistic Analysis? ................. Page 40

3.4.4	 How Does Research Help Reduce Uncertainty? .............. Page 41

3.5	 Inherent Variability in Biological Response ....................... Page 42

3.5.1	 Why Consider Populations and Life-Stages? ................. Page 42
Examination of EPA Risk Assessment Principles and Practices Page v

3.5.2	 Are Certain Populations and Life-Stages Always at


Greater Risk? ......................................... Page 43

3.5.3	 How Are Sensitivities to Toxic Effects Considered? .......... Page 43

3.5.4	 Are Unique or Disproportionate Exposures Considered? ....... Page 45

3.5.5	 Is Human Variability Considered in Occupational Cohorts? ..... Page 47

3.5.6	 What Needs To Be Done To Address the Risk Assessment


of Populations and Life-Stages? .......................... Page 48

3.6	 What Is EPA Doing To Improve Assessment of Uncertainty


and Variability? ............................................. Page 49

3.6.1	 How Can EPA Move Forward Toward Quantitative


Characterization of Uncertainty in Dose-Response? ........... Page 49

3.6.2	 Will EPA Move Toward Integrated Uncertainty Analysis? ...... Page 49

4. CONSIDERING INFORMATION GAPS IN HEALTH ASSESSMENTS: USE OF


DEFAULT AND EXTRAPOLATION ASSUMPTIONS ................... Page 51

4.1	 Default Assumptions ......................................... Page 51

4.1.1	 How Does EPA Use Default Assumptions? ................. Page 51

4.1.2	 Are Default Assumptions Science Policy? .................. Page 52

4.1.3	 Does Any Change Seen in Animals Indicate There Will Be


a Problem for Humans? ................................. Page 53

4.1.4	 Are Benign (Histologically Non-Malignant) Tumors Presumed


To Have Potential To Progress to Malignant Tumors,


and Are They Counted as if They Were Malignant? ........... Page 54

4.1.5	 Should We Expect Target Organ Concordance Between


Effects Observed in Animals and Those Expected in Humans? .. Page 55

4.2	 Extrapolations .............................................. Page 56

4.2.1	 Why Does EPA Use Experiments With Animals To Predict


Effects in Humans? .................................... Page 56

4.2.2	 How Does EPA Adjust for the Differences Between Animals


and Humans? ......................................... Page 57

4.2.3	 Does EPA Use the Most Sensitive Animal Results To Predict


Effects in Humans? .................................... Page 57

4.2.4	 Is an Additional Body Scaling Factor for Children Considered? .. Page 59

4.2.5	 How Are Effects Observed at High Exposures Used To Predict


Responses to Much Lower Exposures? ..................... Page 59

4.2.6	 Why Is the Maximum Tolerated Dose Used? ................ Page 61

4.2.7	 Does the Use of a Maximum Tolerated Dose Affect Cancer


Potency Estimates? .................................... Page 61

4.2.8	 Are the Cancer Risks Estimated by EPA the Expected


Incidence of Cancer for a Given Exposure? ................. Page 62
Page vi Examination of EPA Risk Assessment Principles and Practices 

4.2.9	 How Does EPA Evaluate Incidence Versus Mortality for


Cancer Rates? ......................................... Page 63

4.2.10 If Data Are Only Available for One Route of Exposure,


Does EPA Evaluate Other Routes of Exposure? .............. Page 64

4.3	 Use of Uncertainty Factors ..................................... Page 67

4.3.1	 How Are NOAELs or LOAELs Used for Extrapolations? ...... Page 67

4.3.2	 What Uncertainty Factors Does EPA Use To Reduce the


Experimental NOAEL in Health Assessments? .............. Page 68

4.3.3	 Does EPA Consider the Effects of Combining Several UFs? .... Page 69

4.3.4	 Does EPA Use UFs Derived for Specific Chemicals? .......... Page 71

4.4	 Weight of Evidence .......................................... Page 71

4.4.1	 How Does EPA Consider Weight of Evidence? .............. Page 71

4.4.2	 Does EPA Give More Weight to Positive Animal Studies


Than to Negative Animal Studies? ........................ Page 72

4.4.3	 Does EPA Sometimes Weight Animal Data More Than


Human Data? ......................................... Page 72

4.5	 Toxicity Equivalency Factor Approach ........................... Page 83

4.5.1	 What Are TEFs? ...................................... Page 83

4.5.2	 How Did the TEF Approach Evolve? ...................... Page 85

4.5.3	 Is the Additivity Assumption for TEF Justified? .............. Page 87

4.5.4	 What Is EPA’s Experience With TEFs? .................... Page 87

4.6	 Model Use ................................................. Page 88

4.6.1	 Why Does EPA Use Environmental Models? ................ Page 88

4.6.2	 How Does EPA Approach Environmental Models? ........... Page 89

4.6.3	 What Is EPA Doing To Improve the Use of Environmental


Models? ............................................. Page 89

4.6.4	 How Does EPA Approach Fate and Transport Modeling and


Account for Uncertainty? ................................ Page 92

4.6.5	 Does EPA Consider or Review Existing Data on Environmental


Concentrations? ....................................... Page 94

4.6.6	 Does EPA Consider and Use Case-Specific Data in Models?


(Or: How Does EPA Decide To Modify Default Assumptions


in Later Tiers Using Models?) ............................ Page 95

4.6.7	 How Will EPA Improve Pesticide Fate and Transport


Modeling in the Future? ................................. Page 95

4.6.8	 Does EPA Use Worst-Case Assumptions in Modeling That Are


Unreasonable and Do Not Account for Degradation, Partitioning,


and “Sinks” in the Environment? .......................... Page 96
Examination of EPA Risk Assessment Principles and Practices Page vii


5. SITE- AND CHEMICAL-SPECIFIC ASSESSMENTS ........................ Page 99

5.1	 Overview .................................................. Page 99

5.1.1	 What Are the Superfund Program’s Purpose and Regulatory


Requirements? ....................................... Page 101

5.1.2	 Whom Does the Superfund Program Seek To Protect? The


Reasonable Maximum Exposure Scenario ................. Page 101

5.1.3	 Is the RME Overly Conservative? ........................ Page 102

5.2	 Superfund Risk Assessment Guidance .......................... Page 102

5.2.1	 What Are Some Types of Risk Assessment Under Superfund? . Page 102

5.2.2	 How Does Superfund Use Site-Specific Concentration and


Maximum Concentration Data in Risk Assessment? ......... Page 103

5.2.3	 Has the Superfund Guidance Been Updated and Externally


Peer Reviewed Since It Was First Released in 1989? ......... Page 104

5.3	 Default Assumptions and Site-Specific Information ................ Page 105

5.3.1	 What Is the Role of Default Assumptions and Site-Specific


Information in Superfund Risk Assessments? ............... Page 105

5.3.2	 What Is the Drinking Water Consumption Rate? ............ Page 106

5.3.3	 What Is the Inhalation Rate? ............................ Page 106

5.3.4	 What Is the Exposure Duration for Residences? ............. Page 106

5.3.5	 What Is the Exposure Duration for Workers? ............... Page 107

5.3.6	 What Is the Ingestion Rate for Construction Workers? ........ Page 108

5.3.7	 What Are the Incidental Soil Ingestion Rates for Children? .... Page 108

5.4	 Site-Specific Information ..................................... Page 109

5.4.1	 When Does EPA Use Site-Specific Information? ............ Page 109

5.4.2	 How Is Bioavailability Addressed? ....................... Page 110

5.4.3	 How Are Fish Ingestion Rates Evaluated? ................. Page 111

5.5	 Other Factors in Superfund Assessments ......................... Page 115

5.5.1	 What Is the Role of Stakeholders in Superfund Assessments? .. Page 115

5.5.2	 How Are Superfund Assessments Factored in Superfund Risk


Management Decisions? ............................... Page 116

5.6	 Superfund Site-Specific Conclusions ........................... Page 119

5.7	 Chemical Mixtures Risk Assessment Methods and Practice .......... Page 120

5.7.1	 What Are the Issues Regarding Chemical Mixture Site


Assessments? ........................................ Page 120

5.7.2	 What Are Dose Addition and Response Addition? ........... Page 121

5.7.3	 How Are Dose Addition and Response Addition Applied


in Practice? .......................................... Page 123

5.7.4	 Are Interaction Data Used in Mixture Risk Assessment? ...... Page 124

5.7.5	 Are Data on Whole Mixtures Used in Risk Assessment? ...... Page 124

5.7.6	 What Other EPA Applications Exist for Mixtures Risk


Assessment? ......................................... Page 125
Page viii Examination of EPA Risk Assessment Principles and Practices 

6. ECOLOGICAL ASSESSMENT ......................................... Page 127

6.1	 Overview ................................................. Page 127

6.1.1	 What Is the EPA Ecological Risk Assessment Approach? ..... Page 128

6.2	 Organism-Level Versus Population-Level Ecological


Risk Assessments ........................................... Page 129

6.3	 Conservatism and Ecological Risk Assessments ................... Page 131

6.3.1	 How Is Conservatism Addressed in Ecological Risk


Assessments? ........................................ Page 131

6.3.2	 How Are Toxicity Reference Values Developed in Ecological


Risk Assessments? .................................... Page 131

6.3.3	 How Do Screening-Level and Definitive (Baseline)


Assessments Differ? ................................... Page 132

6.3.4	 How Is Site-Specific Information Used To Derive Exposure


Concentrations in Ecological Risk Assessments? ............ Page 132

6.4	 Water Quality Criteria ....................................... Page 133

6.4.1	 What Is the Agency Standard Methodology for Deriving


Aquatic Life Water Quality Criteria? ...................... Page 133

6.4.2	 What Is the GLI Tier II Methodology? .................... Page 134

6.4.3	 Was the GLI Tier II Methodology Made Available for Public


Comment or Vetted Through Other Public Processes? ........ Page 134

6.5	 Uncertainty and Pesticide Ecological Assessment .................. Page 135

6.5.1	 How Does EPA Address Uncertainty Analysis in Pesticide


Ecological Assessments? ............................... Page 135

6.5.2	 How Does OPP Address Rodenticide Use Patterns? .......... Page 136

6.6	 Summary and Conclusions .................................... Page 139

7. SUMMARY AND RECOMMENDATIONS ............................... Page 141

7.1	 Summary ................................................. Page 141

7.2	 Several Current EPA Activities That Will Enhance EPA Risk


Assessment Principles and Practices ............................ Page 142

7.3	 Recommendations for Further Improvements to Risk Assessment


Principles and Practices ...................................... Page 143

LIST OF USEFUL ABBREVIATIONS AND ACRONYMS ..................... Page 148

GENERAL REFERENCES ............................................... Page 154

REFERENCES OF EPA RISK ASSESSMENT GUIDELINES ................... Page 180

ADDITIONAL USEFUL WEB SITES ...................................... Page 182
Examination of EPA Risk Assessment Principles and Practices Page ix 

ACKNOWLEDGM ENTS 

Many people worked on and contributed to the examination of risk assessment principles 

and practices that ultimately resulted in this document. Many EPA employees from the Agency’s 

Offices and Regions provided input and we would like to specifically acknowledge the efforts 

and contributions made by the following individuals: 

Daniel Axelrad (OPEI)


Timothy Benner (ORD)


Michael Beringer (Region 7)


George Bollweg (Region 5)


David Bussard (ORD)


Michael Callahan (Region 6)


Catherine Campbell (OPEI)


Laurel Celeste (OGC)


Weihsueh Chiu (ORD)


Patricia Cirone (Region 10)


Jenny Craig (OAR)


Larry Cupitt (ORD)


Julie Damon (ASPH/EPA Fellow)


Kerry Dearfield (OSA) - Chair, Task Force


Vicki Dellarco (OPP)


Lynn Flowers (ORD)


Brenda Foos (OCHP)


Stiven Foster (ORD)


Gary Foureman (ORD)


Jack Fowle (ORD)


Herman Gibb (ORD)


Andrew Gillespie (detail from Dept of


 Agriculture) 

David Guinnup (OAR) 

Karen Hammerstrom (ORD) 

Belinda Hawkins (OEI) 

Tala Henry (OW) 

Oscar Hernandez (OPPT) 

Richard Hertzberg (ORD) 

Robert Hetes (ORD) - Chapter 2 lead 

James Hetrick (OPP) 

Ann Johnson (OPEI) 

Myra Karstadt (OPPT) 

Aparna Koppikar (ORD)


Michael Kravitz (ORD)


Stephen Kroner (OSWER)


Patricia Lafornara (OEI)


Elizabeth Leovey (OPP)


John Lipscomb (ORD)


Mary Manibusan (ORD)


Carl Mazza (OAR)


Robert McGaughy (ORD)


Michael Metzger (OPP)


Jayne Michaud (OSWER) - Chapter 5 lead


Amy Mills (ORD)


Kenneth Mitchell (OAR)


Jacqueline Moya (ORD)


Debbie Newberry (OSWER)


Susan Norton (ORD)


Marian Olsen (Region 2)


Jennifer Orme-Zavaleta (ORD)


Nicole Paquette (OEI)


Pasky Pascual (ORD)


Resha Putzrath (ORD) - Chapter 4 lead


Mary Reiley (OW)


Susan Reith (ORD)


Glenn Rice (ORD)


Charles Ris (ORD)


Bruce Rodan (ORD)


John Schaum (ORD)


Rita Schoeny (OW)


Brad Schultz (Region 5) - Chapter 3 lead


Cheryl Scott (ORD)


Suhair Shallal (SAB)


Michael Slimak (ORD) - Chapter 6 lead


Daniel Stralka (Region 9)

Page x Examination of EPA Risk Assessment Principles and Practices 

Glenn Suter (ORD) Douglas Urban (OPP) 

Jeffrey Swartout (ORD) Winona Victery (Region 9) 

Christina Swartz (OPP) Donn Viviani (OPEI) 

Linda Teuschler (ORD) Katie Warwick (ORD) 

Nelson Thurman (OPP) 
Examination of EPA Risk Assessment Principles and Practices Page 1 

U.S. Environmental Protection Agency

AN EXAMINATION OF EPA RISK ASSESSMENT


PRINCIPLES AND PRACTICES

Page 2 Examination of EPA Risk Assessment Principles and Practices 

1. INTRODUCTION TO EPA RISK ASSESSMENT 

1.1 Overview 

1.1.1 What Is Risk Assessment? 

The most common basic definition of risk assessment used within the U.S. 

Environmental Protection Agency (EPA) is paraphrased from the 1983 report Risk Assessment in 

the Federal Government: Managing the Process (NRC, 1983), by the National Academy of 

Sciences’ (NAS’s) National Research Council (NRC): 

Risk assessment is a process in which information is analyzed to determine if an 

environmental hazard might cause harm to exposed persons and ecosystems. 

This process is highly interdisciplinary in that it draws from such diverse fields as 

biology, toxicology, ecology, engineering, geology, statistics, and the social sciences to create a 

rational framework for evaluating environmental hazards. While this definition has been 

somewhat enhanced and elaborated upon through subsequent NAS writings, it still basically 

describes risk assessment as it is performed within EPA. EPA uses risk assessment as a tool to 

integrate exposure and health effects or ecological effects information into a characterization of 

the potential for health hazards in humans or other hazards to our environment. 

1.1.2 Why Does EPA Conduct Risk Assessments? 

The mission of the EPA is to protect human health and to safeguard the natural 

environment — air, water, and land — upon which life depends. EPA fulfills this mission by, 

among other things, developing and enforcing regulations that implement environmental laws 

enacted by Congress. The implementation of environmental laws may include grants and other 

financial assistance to state and tribal governments carrying out environmental programs 

approved, authorized, or delegated by EPA. 

Determining environmental standards, policies, guidelines, regulations, and actions 

requires making decisions. Environmental decision making is often a controversial process 

involving the interplay among many forces: science, social and economic factors, political 

considerations, technological feasibility, and statutory requirements. There are often conflicting 

interests regarding these various forces than can have a bearing on environmental decisions. 

Setting an environmental standard that is too lax may threaten public health, while a standard that 

is unnecessarily stringent may impose a significant marginal economic cost for small marginal 

gain. Environmental decisions are often time-sensitive, for example when public health is 

known or suspected to be at risk. The decisions must frequently be made with incomplete or 

imperfect information and many times under the additional pressure of heightened public 
Examination of EPA Risk Assessment Principles and Practices Page 3 

scrutiny and concern. And, once made, the decisions are often challenged in court and subject to 

high levels of public and scientific scrutiny. 

EPA conducts risk assessment to provide the best possible scientific characterization of 

risks based on a rigorous analysis of available information and knowledge — that is, a 

description of the nature and magnitude of the risk, an interpretation of the adversity of the risk, a 

summary of the confidence or reliability of the information available to describe the risk, areas 

where information is uncertain or lacking completely, and documentation of all of the evidence 

supporting the characterization of the risk. EPA then incorporates this risk characterization with 

all of the other relevant information — social, economic, political, and regulatory — in making 

decisions (policies, regulations) about how to manage the risk. Risk assessment, therefore, 

informs decision makers about the science implications of the risk in question. Risk assessments 

that meet their objectives can help guide risk managers to decisions that mitigate environmental 

risks at the lowest possible cost and which will stand up if challenged in the courts. 

1.1.3 How Does EPA Use Risk Assessments in Decision Making? 

The primary purpose of a risk assessment is to inform the risk manager’s decision making 

process. The primary purpose of a risk assessment is not to make or recommend any particular 

decisions; rather, it gives the risk manager information to consider along with other pertinent 

information. EPA uses risk assessment as a key source of scientific information for making 

good, sound decisions about managing risks to human health and the environment. Examples of 

such decisions include deciding permissible release levels of toxic chemicals, granting permits 

for hazardous waste treatment operations, and selecting methods for remediating Superfund sites. 

The use of credible science in risk assessment helps make and support risk management 

decisions, but it is not the only factor that the risk manager considers. It is generally recognized 

— by the science community, by the regulatory community, and by the courts — that it is 

important to consider other factors along with the science when making decisions about risk 

management. In some regulations, the consideration of other factors is mandated (e.g., costs). 

Some of these other factors include: 

a. Economic factors — the costs and benefits of risks and risk mitigation 

alternatives. 

b. Laws and legal decisions — the framework that prohibits or requires some 

actions. 

c. Social factors — attributes of individuals or populations that may affect their 

susceptibility to risks from a particular stressor. 
Page 4 Examination of EPA Risk Assessment Principles and Practices 

d. Technological factors — the feasibility, impact, and range of risk management 

options 

e. Political factors — interactions among and between different branches and levels 

of government and the citizens they represent. 

f. Public factors — the attitudes and values of individuals and societies with respect 

to environmental quality, environmental risk, and risk management. 

1.1.4 What Is Some Historical Perspective Relevant to EPA Risk Assessment 

Practices? 

EPA was involved with risk assessment practices since EPA’s early days, although risk 

assessment per se was not a formally recognized process then. EPA completed its first risk 

assessment document in December 1975: Quantitative Risk Assessment for Community Exposure 

to Vinyl Chloride (Kuzmack and McGaughy, 1975). The next significant document appeared in 

1976: Interim Procedures and Guidelines for Health Risk and Economic Impact Assessments of 

Suspected Carcinogens (Train, 1976). The preamble of this document, signed by the 

Administrator, signaled the Agency’s intent that “rigorous assessments of health risk and 

economic impact will be undertaken as part of the regulatory process.” A general framework 

described a process to be followed in analyzing cancer risks of pesticides, and the document 

recommended that the health data be analyzed independently of the economic impact analysis. 

Later, in 1980, EPA announced the availability of water quality criteria documents for 64 

contaminants (USEPA, 1980). This was the first application of quantitative procedures 

developed by EPA to a large number of carcinogens, and the first EPA document describing 

quantitative procedures used in risk assessment. 

Then in 1983, the NAS published Risk Assessment in the Federal Government: Managing 

the Process (NRC, 1983; commonly referred to as the “Red Book”). EPA has integrated the 

principles of risk assessment from this groundbreaking report into its practices to this day. The 

following year, EPA published Risk Assessment and Management: Framework for Decision 

Making (USEPA, 1984), which emphasizes making the risk assessment process transparent, 

describing the assessment’s strengths and weaknesses more fully, and providing plausible 

alternatives within the assessment. 

Shortly after the publication of the Red Book, EPA began issuing a series of guidelines 

for conducting risk assessments (e.g., in 1986 for cancer, mutagenicity, chemical mixtures, 

developmental toxicology, and in 1992 for estimating exposures). Although EPA efforts focused 

initially on human health risk assessment, the basic model was adapted to ecological risk 

assessment in the 1990s to deal with a broad array of environmental risk assessments in which 

human health impacts are not directly at issue. EPA continues to make a substantial investment 
Examination of EPA Risk Assessment Principles and Practices Page 5 

in advancing the science and application of risk assessment through continual updates of these 

guidelines and the development of newer guidelines as needed. Refer to the section in the 

References set aside for the listing of these EPA guidelines. 

Over time, the NAS expanded on its risk assessment principles in a series of subsequent 

reports, including Pesticides in the Diets of Infants and Children (NRC, 1993), Science and 

Judgment in Risk Assessment (NRC, 1994; also known as the “Blue Book”), and Understanding 

Risk: Informing Decisions in a Democratic Society (NRC, 1996). For example, the NAS places 

equal emphasis on fully characterizing the scope, uncertainties, limitations, and strengths of the 

assessment and on the social dimensions of interacting with decision makers and other users of 

the assessment in an iterative, analytic-deliberative process. The purpose of this process is to 

ensure that the assessments meet the intended objectives and are understandable. EPA risk 

assessment practices have evolved over time along with this progression of thought, and in many 

cases helped drive the evolution of thinking on risk assessment. 

In 1995, EPA updated and issued the current Agency-wide Risk Characterization Policy 

(USEPA, 1995a). The Policy calls for all risk assessments performed at EPA to include a risk 

characterization to ensure that the risk assessment process is transparent; it also emphasizes that 

risk assessments be clear, reasonable, and consistent with other risk assessments of similar scope 

prepared by programs across the Agency. Effective risk characterization is achieved through 

transparency in the risk assessment process and clarity, consistency, and reasonableness of the 

risk assessment product — TCCR. EPA’s Risk Characterization Handbook (USEPA, 2000a) 

was developed to implement the Risk Characterization Policy. 

The Congressional/Presidential Commission on Risk Assessment and Risk Management 

(CRARM) was created by the Clean Air Act Amendments of 1990 and formed in 1994. Its 

mandate was to make a full investigation of the policy implications and appropriate uses of risk 

assessment and risk management in regulatory programs, under various federal laws, designed to 

prevent cancer and other chronic health effects that may result from exposure to hazardous 

substances. More specifically, its mandate was to provide guidance on how to deal with residual 

emissions from Section 112 hazardous air pollutants (HAPs) after technology-based controls 

have been placed on stationary sources of air pollutants. In 1997, the Commission published its 

report in two volumes (CRARM, 1997a; CRARM, 1997b). These discussed the importance of 

better understanding and quantification of risks, as well as the importance of evaluating strategies 

to reduce human and ecological risks. 

EPA’s risk assessment principles and practices build on our own risk assessment 

guidances and policies — such as the Risk Characterization Policy; Guidance for Cumulative 

Assessment, Part 1: Planning and Scoping (USEPA, 1997a); the Risk Assessment Guidance for 

Superfund, or RAGS (USEPA, 1989a, and subsequent updates); EPA’s Information Quality 

Guidelines (USEPA, 2002a); and A Summary of General Assessment Factors for Evaluating the 
Page 6 Examination of EPA Risk Assessment Principles and Practices 

Quality of Scientific and Technical Information (USEPA, 2003a) — as well as the NAS, the 

CRARM, and others’ concepts. It is understood that risk assessment provides important 

information about the nature, magnitude, and likelihood of possible environmental risks to 

inform decisions — principles that evolved out of these many efforts. 

1.2	 EPA Process for Evaluation of Risk Assessment Principles and Practices 

1.2.1	 Why Are We Conducting an Evaluation of Our Risk Assessment Principles 

and Practices Now? 

EPA constantly evaluates its risk assessment principles and practices, mostly via a 

gradual refinement of particular practices that may not be overtly visible to the public. There are 

times when EPA takes a concentrated, focused approach: for example, when revising a major 

risk assessment guideline such as for cancer assessment. EPA conducts a wider, general review 

of its risk assessment principles and practices occasionally to help strengthen core values and 

increase its ability to make better decisions. 

In early 2002, the position of the EPA Science Advisor was established. The Science 

Advisor’s overarching responsibility is to coordinate and oversee the scientific activities of the 

program and regional offices at EPA. Part of this responsibility is to ensure the best use of 

science at the Agency and in its decisions. At the Science Advisor’s request, EPA staff began 

looking at the Agency’s risk assessment practices and training with an eye to update them. When 

the Office of Management and Budget (OMB) solicited comments on risk assessment practices 

across the federal government (see section 1.2.4), we (EPA staff) took this as an opportunity to 

concentrate on a wider review to evaluate current risk assessment practices across programs and 

regions. 

1.2.2	 What Is the Purpose and Intent of This Staff Paper? 

This staff paper was developed to give the EPA scientific and technical professional staff 

an opportunity to present what we (EPA staff) believe are the current EPA risk assessment 

principles and practices. The practices are presented in the context of the public comments 

submitted to OMB. The paper’s purpose is first to open a dialogue among EPA risk assessors 

and risk managers about Agency risk assessment practices. Then, as we engage the public, we 

will continue the dialogue about how we can move forward together to clarify and, where 

appropriate, strengthen our risk assessment practices. 

The staff paper is intended for a wide audience of people who are very familiar with risk 

assessment principles and practices -- risk assessors and risk managers within the EPA as well as 

those outside EPA with knowledge of risk assessment practices. The discussion contained here 

is not meant to be a primer on risk assessment or an introduction for those not very familiar with 
Examination of EPA Risk Assessment Principles and Practices Page 7 

these practices. Since the comments address complex and sometimes subtle nuances of risk 

assessment, the staff paper attempts to deal with those comments and concepts at that level of 

discussion. 

1.2.3	 Through What Process Is EPA Conducting This Evaluation? 

Specifically for this evaluation, on June 17, 2003, an Agency-wide memorandum was 

issued from three EPA senior managers (Jessica Furey, Associate Administrator for OPEI; Paul 

Gilman, EPA Science Advisor and Assistant Administrator for ORD; and Stephen Johnson, 

Assistant Administrator for OPPTS) to start this evaluation. The memorandum called for the 

establishment of an Agency-wide workgroup, the Risk Assessment Task Force, to review risk 

assessment principles and practices at the Agency. The Risk Assessment Task Force focused on 

the practices, assumptions, defaults and principles identified in the comments sent to OMB (see 

section 1.2.3 below), as well as issues identified from within EPA. Generation of this report on 

the initial analyses of the Task Force (EPA staff) is the first step in a multi-step process. In the 

future, EPA expects to communicate with stakeholders about the results of this staff evaluation 

and give them opportunities for dialogue, in order to understand their concerns. 

1.2.4	 How Does the 2003 OMB Draft Report to Congress Relate to Our 

Evaluation? 

On February 3, 2003, OMB published in the Federal Register (FR) a request for 

comments on its Draft 2003 Report to Congress on the Costs and Benefits of Federal 

Regulations (USOMB, 2003a). In this FR notice, OMB also sought the public’s views on a 

number of important issues pertaining to the practice of risk assessment. OMB received many 

public comments and these were passed onto EPA. Our effort takes advantage of this 

information as we review risk assessment principles and practices at the Agency. 

OMB issued its final report to Congress on regulatory policy in September 2003 

(USOMB, 2003b) based on the comments it received on the February draft report. The final 

report presents findings on major federal rulemakings finalized over the previous 10 years, 

specific regulatory reforms, guidance on regulatory analysis, homeland security proposals, and 

Agency consultations with states and local governments. The report also deals with the concept 

of precaution in U.S. approaches to risk assessment and management. OMB concludes that 

precaution plays an important role in risk assessment and risk management, but precaution, 

coupled with objective scientific analysis, needs to be applied wisely on a case-by-case basis. 
Page 8 Examination of EPA Risk Assessment Principles and Practices 

1.2.5	 What Is the General Nature of the Comments on EPA Risk Assessment 

Practices Submitted to OMB? 

The vast majority of the comments can be generally summarized from the American 

Chemistry Council (ACC) submission. ACC’s basic recommendations are not radical or 

particularly new. They reiterate three points that can be found in the 1991 Executive Office of 

the President document Regulatory Program of the United States Government (EOP, 1991): 

a) Risk assessments should not continue an unwarranted reliance on “conservative 

(worst-case) assumptions” that distort the outcomes of the risk assessment, 

“yielding estimates that may overstate likely risks by several orders of 

magnitude.” 

b) Risk assessments should “acknowledge the presence of considerable uncertainty” 

and present the extent to which conservative assumptions may overstate likely 

risks. 

c) EPA risk assessments must not “intermingle important policy judgments within 

the scientific assessment of risk.” Rather, the “choice of an appropriate margin of 

safety should remain the province of responsible risk-management officials, and 

should not be preempted through biased risk assessments.” 

On the other hand, various other comments submitted suggest that EPA risk assessments 

do not fully address all risks. Generally, these latter comments relate to issues of cumulative and 

aggregate risk. They state that EPA risk assessments concentrate on single chemical/stressor 

risks, failing to account for multiple chemical/stressor exposures and other factors as life-stage, 

lifestyle, and increased susceptibility of certain exposed populations. This practice would tend to 

underestimate risk in real-world scenarios. 

1.3	 Organization of This Document 

1.3.1	 How Are the General Comments Addressed? 

This document examines EPA risk assessment principles and practices in light of the 

general comments outlined in the preceding section. The comments focus on issues of 

conservatism in risk assessment (e.g., overstated/understated risks), use of rigid default 

assumptions, poor transparency in the risk assessments we produce, and unacknowledged 

uncertainty. The following chapters are designed to address these general comments (issues). 

Chapter 2 deals with the issue of conservatism. Chapter 3 discusses the nature of uncertainty and 

variability and how EPA deals with these in risk assessment. Chapter 4, on defaults and 

extrapolations, discusses how assumptions can be used when chemical- and/or site-specific data 
Examination of EPA Risk Assessment Principles and Practices Page 9 

are not available or are inadequate to use in performing a risk assessment. The next chapter, 

chapter 5, approaches the use of site- and chemical-specific data. This chapter emphasizes the 

Superfund approaches — many of the specific comments we received were specific to 

Superfund. 

Most of the comments EPA received focused on human health assessment. Many of the 

principles and practices, though, apply to ecological health assessments as well. Chapter 6 

briefly discusses and evaluates our ecological health assessments. 

The last chapter, chapter 7, provides a summary of the Risk Assessment Task Force 

conclusions and recommendations, with transparency in our risk assessment practices as a focus. 

1.3.2	 How Are the Specific Comments and Examples Addressed? 

This document is organized to address the overarching issues of conservatism, 

uncertainty, and transparency. However, many comments focused on specific risk assessment 

principles and practices as well as specific chemical assessments as examples. We will not 

discuss every specific comment in a “one to one” discussion, although there are some instances 

in the following chapters for which specific examples are provided to illustrate a point. While 

we will mention specific chemicals and specific rulemakings in the context of the general issues 

discussion, we will not discuss individual chemical- or site-specific decisions, as these are 

generally of high importance to many, are complex, and (in many cases) are still in active 

interaction with interested stakeholders. 

1.4	 Other Components Impacting Risk Assessment, But Not Addressed in This 

Document 

1.4.1	 What Impact Do the OMB and EPA Information Quality Guidelines Have 

on EPA Risk Assessment Practices? 

The OMB guidelines state that information needs to be “objective, realistic, and 

scientifically balanced” (USOMB, 2002). EPA embraced the OMB guidance when developing 

its Information Quality Guidelines (USEPA, 2002a). These guidelines point to our reliance on 

the extensive use of peer review, the practices found in the various risk assessment guidelines 

and the Risk Characterization Policy, and the Agency-wide use of EPA’s Quality System. These 

EPA practices help ensure that information we use is objective, realistic, and scientifically 

balanced. As we stress, EPA uses scientific peer review to help ensure the quality of the risk 

assessments we generate and to keep the assessments as objective, and as consistent, as possible. 

EPA’s Peer Review Handbook (USEPA, 2000b) stresses that peer review panels be balanced in 

terms of their expertise and biases so that a reasonable and scientifically balanced review results. 

Further, conflicts of interest need to be readily identified. The Risk Characterization Policy 
Page 10 Examination of EPA Risk Assessment Principles and Practices 

encourages presentation of all plausible viewpoints that are realistic and scientifically 

supportable. In 1984, under EPA Order 5360.1, an Agency-wide Quality System was 

established. In 2000, Agency-wide policies were defined in EPA Order 5360.1 A2 and EPA 

Manual 5360 A1, which expanded the policy to accommodate current and evolving needs of the 

Agency. EPA’s Quality System helps ensure that we stay objective in the use of information in 

any assessment we generate (USEPA, 2000c, 2000d). In association with the Information 

Quality Guidelines, we also developed general assessment factors related to quality issues 

regarding scientific and technical information (USEPA, 2003a). One should take these 

assessment factors into consideration as appropriate when evaluating the quality and relevance of 

information, regardless of source. 

Many parties believe that EPA risk assessment practices do not follow this guidance. We 

believe EPA conducts risk assessments consistent with this guidance, as evidenced by some of 

the efforts detailed in the paragraph above. We continually strive to publicly present our risk 

assessment practices and guidance documents and subject them to peer review. Our Risk 

Characterization Policy directs us to consider all scientifically plausible and supportable 

viewpoints, but this information is only part of the full range of information risk managers use to 

make a decision. 

1.4.2 Why Is Peer Review So Important? 

The value of scientific peer review in ensuring the quality of our scientific and technical 

products is critical to EPA and is widely understood and accepted across the Agency. 

Conscientious use of peer review is essential to the credibility of the risk assessments EPA uses 

to support its decisions. Consistent Agency-wide application of peer review has been an EPA 

priority for many years and continues to be so. Since issuing the peer review policy in 1993, 

EPA has continually supported and strengthened the policy and practices of peer review, which 

in turn have supported and strengthened our risk assessment principles and practices. For 

example, peer review gave us an opportunity to seek expert review on many of the issues 

highlighted in this document, such as the reasonableness of assumptions and methodologies used 

in Agency risk assessments. 

A peer review is a documented critical review of a specific scientific or technical work 

product, conducted by qualified individuals (or organizations) who are independent of those who 

performed the work (at a minimum, from a different office) but who are collectively equivalent 

in technical expertise (i.e., peers). EPA’s Peer Review Handbook (2nd Edition; USEPA 2000b) is 

one of the most advanced treatments of peer review for intramural research and 

scientific/technical analysis of any federal agency. By utilizing the practices embodied in the 

Handbook, EPA ensures the credibility and quality of the risk assessments we generate to support 

Agency decisions. 
Examination of EPA Risk Assessment Principles and Practices Page 11 

2. EPA RISK ASSESSMENT AND PUBLIC AND 

ENVIRONMENTAL HEALTH PROTECTION 

2.1 Public and Environmental Health Protection (“Public Health Protection”) 

EPA’s risk assessments are conducted in support of its mission to protect public health 

and the environment. Given the uncertainty, variability, and data gaps encountered when 

conducting any risk assessment, a key objective for EPA's risk assessments is that they avoid 

both underestimation of risk and gross overestimation of risk. 

2.1.1 What Is EPA’s General Approach for Developing Risk Assessments? 

Over the years, practices have been developed to assess risk based on available data and 

information, and EPA has been at the forefront of much of this development. Risk assessment is 

a complex process, requiring the integration of data and information across a broad range of 

activities and disciplines, including source characterization, fate and transport, modeling, 

exposure assessment, and dose-response assessment. EPA seeks to use the available information 

(data) in an objective, realistic, and scientifically balanced way. In each specific assessment, the 

Agency incorporates the relevant data and information to the extent possible (e.g., see chapter 5). 

Where relevant chemical- and/or site-specific data are not available, EPA uses specific default 

assumptions and extrapolations to fill in the data gaps and allow the risk assessment to proceed 

(see chapter 4) so that the Agency can ultimately make the decisions required under its mandates. 

This approach is consistent with the NRC’s recommendation about EPA’s use of defaults (NRC, 

1994). In general, EPA’s default assumptions are based on peer reviewed studies, empirical 

observations, extrapolation from related observations, or scientific theory. 

2.1.2 What Is the “Conservatism” Issue in Terms of Public Health Protection? 

Because of data gaps, as well as uncertainty and variability in the available data, risk 

cannot be known or calculated with absolute certainty. Further, as Hill (1965) noted, a lack of 

certainty or perfect evidence “does not confer upon us a freedom to ignore the knowledge we 

already have, or to postpone the action that it appears to demand at a given time.” Therefore, 

consistent with its mission, EPA risk assessments tend towards protecting public and 

environmental health by preferring an approach that does not underestimate risk in the face of 

uncertainty and variability. In other words, EPA seeks to adequately protect public and 

environmental health by ensuring that risk is not likely to be underestimated. However, because 

there are many views on what “adequate” protection is, some may consider the risk assessment 

that supports a particular protection level to be “too conservative” (i.e., it overestimates risk), 

while others may feel it is “not conservative enough” (i.e., it underestimates risk). This issue 

regarding the appropriate degree of “conservatism” in EPA’s risk assessments has been a concern 
Page 12 Examination of EPA Risk Assessment Principles and Practices 

from the inception of the formal risk assessment process and has been a major part of the 

discussion and comments surrounding risk assessment. 

Even with an optimal cost-benefit solution, in a heterogeneous society, some members of 

the population will bear a disproportionate fraction of the costs while others will enjoy a 

disproportionate fraction of the benefits (Pacala et al., 2003). Thus, inevitably, different 

segments of our society will view EPA’s approach to public health and environmental protection 

with different perspectives. 

The NRC in its 1994 report (NRC, 1994) examined the conservatism issue, particularly as 

it relates to the default assumptions EPA uses. The Committee did not reach consensus on the 

degree to which “conservatism” should play a role in defining defaults. This issue was heavily 

debated to the point where divergent views were offered, one advocating the principle of 

“plausible conservatism” and the other the maximum use of scientific information in selecting 

defaults. In general, EPA uses defaults that guard against underestimating risk while also being 

scientifically plausible given existing uncertainty. EPA’s use of various default assumptions is 

the basis for many of the differences of opinion about its risk assessment practices. This 

concern, identified by the NRC in 1994, continues today as indicated by many of the comments 

submitted to OMB. 

The question of conservatism is heightened by the ambiguous definitions and uses of the 

term “conservatism” by the various concerned parties, including those that feel EPA 

overestimates risk and those that feel the Agency systematically understates risks. Some of the 

various concepts associated with term “conservatism” include prudence versus misestimation, 

conservatism as a response to uncertainty or variability, “level of conservatism,” and “amount of 

conservatism” (NRC, 1994). 

2.1.3 Is Science Policy Utilized Within the Risk Assessment Process? 

Science policy positions and choices are by necessity utilized during the risk assessment 

process. Two major ways in which the risk assessment process uses science policy are described 

below in this section. Note that the utilization of science policy in the risk assessment process is 

not meant to “bury” or “hide” risk management decisions within the risk assessment. The use of 

any policy choice needs to be transparent in a risk assessment. In addition, although science 

policy is utilized in the risk assessment process, it is important to recognize that the policy 

positions themselves are developed outside the risk assessment. These policy positions are 

usually supported by scientific data and/or consensus and ensure that the risk assessment 

proceeds in a way that best serves the needs of the decision maker and the Agency. 

First, there are some basic, fundamental policy positions that frame the risk assessment 

process to ensure that the risk assessments produced are appropriate for a particular decision. 
Examination of EPA Risk Assessment Principles and Practices Page 13 

For example, since EPA is a health and environmental protective agency, EPA’s policy is that 

risk assessments should not knowingly underestimate or grossly overestimate risks. This policy 

position prompts risk assessments to take a more “protective” stance given the underlying 

uncertainty with the risk estimates generated. Another framing policy position is that EPA will 

examine and report on the upper end of a range of risks or exposures when we are not very 

certain about where the particular risk lies. For example, in a screening-level risk assessment for 

hazardous air pollutants (HAPs), the risk assessment starts with a 70-year exposure to assess an 

individual most exposed (see section 2.2.7 for more discussion on this specific example). 

Further, when several parameters are assessed, upper-end values and/or central tendency values 

are generally combined to generate a risk estimate that falls within the higher end of the 

population risk range. Currently, the use of the upper part of a range pertains more often to the 

exposure component of the risk assessment than the hazard/dose-response portion. Many 

comments to EPA suggest that the combining of upper ends leads to unreasonable estimates of 

risk. We generally believe otherwise (e.g., see section 2.1.6), and we feel that this practice 

should be explained clearly in the risk characterization to ensure that risk managers can make an 

appropriate decision consistent with the Agency’s goal of public health and environmental 

protection. These policy positions not only shape the risk assessment process, but are also a 

factor in the decision making process outside the risk assessment. 

Second, default assumptions utilized in any given risk assessment entail science policy 

positions or choices. These science policy choices are more specific than the framing science 

policies, but generally are consistent with the framing policies. For example, a change that is 

considered adverse (i.e., associated with toxicity) in an animal study is assumed to indicate a 

problem for humans unless data demonstrate otherwise. As discussed more fully in chapter 4 

(particularly sections 4.1.1. and 4.1.2), default assumptions are generally supported by scientific 

data and/or scientific consensus. Their use in risk assessments is to allow the risk assessment to 

proceed when chemical- and/or site-specific data are missing or not useful. 

Most importantly, any science policy position or choice used in the risk assessment 

process does not direct the risk assessment itself toward a specific risk management decision, 

e.g., the use of a specific risk estimate. Rather, the risk assessment informs the decision maker 

about the potential risks and uncertainties around the risk estimate(s). These characterized risks 

are then considered in light of the other factors before a decision is made (see section 1.1.3); 

science policy is one of the factors that help a risk manager make a decision. 

2.1.4	 What Impact Does Statutory Language Have on EPA Risk Assessment 

Practices for Public Health Protection? 

The discretionary power afforded to agencies in making regulatory decisions varies 

greatly (OSTP, 1995). EPA faces regulatory, licensing, and other decisions covering a wide 

range of environmental issues and pollutants. These decisions are made within a number of EPA 
Page 14 Examination of EPA Risk Assessment Principles and Practices 

program offices, each responding to a unique mixture of statutes, precedents, and stakeholders. 

Congress establishes legal requirements that generally describe the level of protectiveness that 

EPA regulations must achieve and, infrequently, Congress imposes specific risk assessment 

requirements. In addition, court precedents can affect how EPA considers assessments of risk. 

Given this apparent variety of goals, the constraints within which those goals must be reached, 

and the discretionary powers afforded, it is to be expected that decisions made in one EPA 

program office, for one particular environmental issue, may appear at least on the surface to be 

based on different risk considerations and/or public health protection goals from those used in 

other EPA program offices and for other environmental issues. For example, there may appear to 

be a variation from one EPA risk management decision to another in the way in that public health 

protection goals have been balanced against other considerations such as technological 

feasibility, precedent, and cost. However, overall, EPA generally interprets its statutes to be 

protective of public health and the environment. 

Apparent inconsistencies in risk assessment practices across EPA can stem from 

differences in statutory language. For example, individual statutes identify varying risks to 

evaluate and to protect against (e.g., establish a margin of safety; protect sensitive resources; 

reduce overall risks) and mandate different levels of protection (e.g., protect public welfare; 

prevent unreasonable risk; reduce overall risks; function without adverse effects). Examples 

among major EPA program offices illustrate some of the different Congressional mandates 

regarding risk assessment and risk management practices: 

a)	 In the case of threshold effects ... an additional ten-fold margin of safety for the 

pesticide chemical residue shall be applied for infants and children ... (OPPTS; 

FFDCA §408 (b)(2)(C)) 

b)	 The Administrator shall, in a document made available to the public in support of 

a regulation promulgated under this section, specify, to the extent practicable: 

1) 	 Each population addressed by any estimate of public health effects; 

2) 	 The expected risk or central estimate of risk for the specific populations; 

3) 	 Each appropriate upper-bound or lower-bound estimate of risk ... (OW; 

SDWA § 300g-1 (b)(3)) 

c)	 The Administrator shall ... [add] pollutants which present, or may present, 

through inhalation or other routes of exposure, a threat of adverse human health 

effects...or adverse environmental effects through ambient concentrations, 

bioaccumulation, deposition, or otherwise but not including releases subject to 
Examination of EPA Risk Assessment Principles and Practices Page 15 

regulation under subsection (r) of this section as a result of emissions to air... 

(OAR; CAA §112(b)(2)) 

d)	 ... Provide an ample margin of safety to protect public health or to prevent an 

adverse environmental effect (OAR; CAA §112(f)). 

Similarly, individual statutory requirements regarding the appropriate level of protection 

can have a significant impact on the focus (the purpose and scope) of a risk assessment, which 

can lead to the appearance of inconsistency in risk assessment practices. Such requirements vary 

across Agency programs; for example: 

a)	 ... To assure chemical substances and mixtures do not present an unreasonable 

risk of injury to health or the environment (OPPTS; TSCA §2(b)(3)). 

b)	 ... Function without unreasonable and adverse effects on human health and the 

environment (OPPTS; FIFRA §3). 

c)	 ... Necessary to protect human health and the environment (OSWER; RCRA 

§3005 as amended). 

d)	 ... Provide the basis for the development of protective exposure levels (OSWER; 

NCP §300.430(d)). 

e)	 ... Adequate to protect public health and the environment from any reasonably 

anticipated adverse effects (OW; CWA §405(d)(2)(D)). 

Even the statutory language used for different statutes administered within one major office, 

EPA’s Office of Air and Radiation (OAR), shows differences: 

a)	 ... Protect public health with an adequate margin of safety (OAR; CAA §109). 

b)	 ... Provide an ample margin of safety to protect public health or to prevent an 

adverse environmental effect (OAR; CAA §112(f)). 

c)	 ... Protect the public welfare from any known or anticipated adverse effects 

(OAR; CAA §109). 

d)	 ... [Not] cause or contribute to an unreasonable risk to public health, welfare, or 

safety (OAR; CAA §202(a)(4)). 
Page 16 Examination of EPA Risk Assessment Principles and Practices 

e)	 ... Protect sensitive and critically sensitive aquatic and terrestrial resources (OAR; 

CAAA §404 (Appendix B)). 

f)	 ... Reduce overall risks to human health and the environment (OAR; Title VI of 

CAA). 

g)	 ... Actions to mitigate environmental and health risks (OAR; SARA Title IV). 

For example, within OAR’s National Ambient Air Quality Standards Program, the 

primary standards are health-based and do not consider costs or technological feasibility. The 

secondary standards consider the impacts of pollutants on human economic well-being such as 

visibility, agricultural productivity, and ecological impacts. Under the Clean Water Act, EPA’s 

Office of Water (OW) publishes ambient water quality criteria based on protecting human health; 

these risk assessments do not consider the cost or technological feasibility of meeting these 

criteria. Under the Safe Drinking Water Act, OW conducts risk assessments to determine nonenforceable Maximum contaminant level goals (MCLGs). OW then sets enforceable Maximum 

contaminant levels (MCLs) as close as technically feasible to the MCLGs after taking costs into 

consideration. 

In its report on chemical risk assessment, the General Accounting Office, or GAO, 

(USGAO, 2001) recognized how contextual differences in statutes may affect the focus of risk 

assessments. By taking an overall public health protective stance (as a response to uncertainty 

and variability), EPA’s approach to risk assessment takes into account the variety of language 

found in the various statutes and allows some overall consistency in the Agency’s risk 

assessment practices, while allowing specific EPA offices to follow their particular mandates. 

2.1.5	 Does EPA Take a Reasonable Approach to Public Health Protection? 

Since uncertainty and variability are present in risk assessments, EPA usually 

incorporates a “high-end” hazard and/or exposure level in order to ensure an adequate margin of 

safety for most of the potentially exposed, susceptible population, or ecosystem. EPA’s high-end 

levels are around 90% and above — a reasonable approach that is consistent with the NRC 

discussion (NRC, 1994). This policy choice is consistent with EPA’s legislative mandates (e.g., 

adequate margin of safety, see section 2.1.4). Even with a high-end value, there will be exposed 

people or environments at greater risk and at lower risk. In addition to the high-end values, EPA 

programs typically estimate central tendency values for risk managers to evaluate. This provides 

a reasonable sense of the range of risk that usually lies on the actual distribution. 

When EPA has usable data which show that protection at a higher level (e.g., 99%) or 

lower level (e.g., 90% or lower) is appropriate, then managers may use the information to make 

the risk management decision. For example, exposure data used for risk assessment are 
Examination of EPA Risk Assessment Principles and Practices Page 17 

frequently based on studies, such as controlled field trials for pesticides, that reflect high-end 

exposures. Usually, the relationship between these levels and the actual levels to which people 

are exposed is not known. This leads to uncertainty in estimates about the percentile of the 

population exposed in a given assessment. Therefore, the risk management decision regarding 

the percentile at which to regulate needs to consider how close the exposure data are likely to 

approximate actual exposures in the population being considered. Typically, percentiles chosen 

for pesticides regulation can range from possibly the 70th or 80th percentile in situations (e.g., 

occupational) where it is known that use of the available data will clearly overestimate exposure 

to the 99.9th percentile when the data are likely to closely approximate the population exposures 

(e.g., specific dietary exposures). 

Whatever hazard and exposure values are used to estimate risk, it is important to be 

transparent in characterizing the range of possible risks. For example, if we identify certain 

populations that will be at greater risk than the high end we propose, then it is important to 

highlight these populations so decision makers can make appropriate decisions regarding their 

possible risk. The risk assessment should not make that decision; it should characterize that risk 

and who or how much is at risk. Consequently, managers may decide on a greater level of action 

in a certain locale, if conditions warrant, than may be appropriate elsewhere. On the other hand, 

risk assessments should also characterize the range of risk, including the lower levels of risk. 

EPA’s policy is to characterize the range of risk and to also highlight exceptionally susceptible 

populations (USEPA, 1995a, 2000a). 

While the need for a full characterization of risk is stated in EPA’s policies and guidance, 

the actual characterization in risk assessments may not be so explicit. Therefore, we may need to 

more consistently characterize the range of risk and highlight susceptible populations in our risk 

assessments. This greater transparency will help make the reasonableness of the estimated risk 

for the exposed population or ecosystem more understandable. Related to this, the NRC pointed 

out in 1994 that there is little empirical evidence to suggest that EPA’s potency, exposure, or risk 

estimates are markedly higher than estimates embodying a reasonable degree of prudence (i.e., 

the conventional benchmarks of the 95th or 99th percentiles that statisticians use) (NRC, 1994). 

2.1.6 What Happens When Default Assumptions Are Combined? 

Some comments assert that EPA has so overemphasized conservatism that most risk 

estimates are alarmingly false, meaningless, and unscientific. It should be noted that the use of 

default assumptions does not render the process or results non-scientific. The basis of the 

argument that EPA’s risk estimates are alarmingly false suggests that combining several values 

(e.g., use of values at a default level of 95%) results in excessive overestimates of risk. For 

example, it is implied that combining two 95th percentile defaults results in an estimate above the 

99th percentile, that combining three 95th percentile defaults results in an estimate above the 99.9th 
Page 18 Examination of EPA Risk Assessment Principles and Practices 

percentile, and so forth. Suggestions have also been made that using the mean instead of the 95th 

percentile in these types of calculations would result in a less conservative risk estimate. 

However, the final distribution in these types of calculations is influenced differently by 

the different inputs, and just multiplying the numbers as implied above will not lead to the 

answers above. In estimating exposure or risk using percentiles, how much “influence” using a 

particular 95th percentile value has in the ultimate exposure or risk calculation depends on more 

than just its position in the percentile rankings. It depends also on the variability of the data 

within the distribution for the input factors, the shape of the input distributions, and even the 

number of data points. If all the input variables show the same variability, shape, etc., then the 

above reasoning about compounding values is true. This is rarely the case in actual situations, 

however. 

For illustrative purposes, consider a case where the calculated exposure is a product of 

several input factors. If most of these factors have little variability (e.g., the average lifetime of 

70 years, or the volume of air breathed per day varies within plus or minus 25% for almost all the 

population subgroups) and the concentration of a chemical that the population is exposed to 

varies by three orders of magnitude, then the resulting percentile position of the calculated 

exposure within the resulting exposure distribution will be much more influenced by which value 

is selected from the input distribution of concentrations than it will be for which value is selected 

for the other factors. Selecting the mean value for the concentration input value and 95th 

percentile values for the others will result in a calculated exposure that is much closer to the 

mean of the resulting distribution than the 95th percentile (or higher), because the resulting 

distribution is heavily influenced by the concentration input. Conversely, selecting the 95th 

percentile from the concentration input distribution and the means of the others will result in a 

calculated exposure that is close to the 95th percentile of the resulting exposure distributions. 

Consequently, in the cases where all the input distributions are not the same in variability, where 

the final estimate falls on the combined distribution depends on which input variable is selected 

as 95th percentile. 

In fact, the example above is similar to the RME (reasonable maximum exposure) 

calculated in Superfund assessments. Typically, the concentrations in environmental media are 

highly variable and the other parameters are less variable. Selecting the mean of the 

concentration distribution, while setting one or more of the other parameters at the 95th percentile 

value, usually results in an estimate that is reasonably conservative (with the acknowledgment 

that each case, due to its unique data set, will be different). 

An additional factor within the RME calculation is that the 95% upper confidence limit 

(UCL) on the mean of the concentrations is used instead of the mean itself. Using the mean 

concentration for an area with contaminated soil is in effect setting a scenario where it is equally 

probable that the exposed individual will be at any given location in the area. When the site has 
Examination of EPA Risk Assessment Principles and Practices Page 19 

“attractive nuisances” with higher concentrations, the risk assessor factors this into the 

assessment. The coverage of the data set within the area affects the confidence in the mean 

calculated from the data. For example, confidence in calculating the true mean for a one-acre site 

with three data points will be much less than for a one-acre site with 300 data points. The 95% 

UCL allows the assessor to have some confidence that the mean is not being underestimated for a 

relatively sparse data set. On the other hand, for a robust data set, the 95% UCL will be quite 

close to the mean itself, so it does not introduce appreciable additional conservatism into the 

estimate of exposure. Further, the upper confidence limit on the mean is not the same as the 95th 

percentile of observed concentrations. 

For many other exposure calculations that EPA performs, the Agency’s 1992 Guidelines 

for Exposure Assessment (USEPA, 1992a) provide specific guidance on how to construct 

estimates using distributions of contributing factors, taking into consideration that the resulting 

exposure estimate will be more sensitive to some factors than others. These guidelines suggest 

that when exposure data or probabilistic simulations are not available, an exposure estimate that 

lies between the 90th percentile and the maximum exposure in the exposed population be 

constructed “by using maximum or near-maximum values for one or more of the most sensitive 

variables, leaving others at their mean values” (USEPA, 1992a). 

Nor is just using the mean instead of the 95th percentile necessarily less “conservative.” 

A percentile distribution, such as the collection of data associated with soil concentrations on a 

specific site, is an ordered series of data values. The values for the data may be such that the 95th 

percentile value may be higher or lower than the mean. While the mean of a data set is most 

often below the 95th percentile, consider a data set of 100 soil samples in which 97 samples are at 

the background level, 1 part per million (ppm), for a given pollutant, while the remaining 

samples are at 50 ppm, 100 ppm, and 1,700 ppm. The 95th percentile of this data set is 1 ppm, 

which is actually numerically indistinguishable from the minimum value; the mean, 1.947 ppm, 

lies between the 96th and 97th percentile. Consequently, use of the 95th percentile value in a 

calculation instead of the mean does not necessarily make the calculation more — or less — 

conservative. While the mean does indeed fall below the 95th percentile for the vast majority of 

cases EPA sees, the fact remains that mathematical manipulation of percentile data and selection 

of values from percentile distributions for purposes of estimating resulting exposures in a given 

scenario needs to be considered on an individual basis to determine the degree of “conservatism.” 

Risk estimates often involve adding risks from various chemicals. This can be done by 

adding doses for chemicals with the same mode of action (the approach used by EPA’s Office of 

Pesticide Programs, or OPP), or it can be approximated by adding, for example, the resulting 

single-chemical cancer risks for various chemicals (the approach EPA uses in permitting and in 

its Superfund program). Adding individual chemical risks to estimate a combined, cumulative 

risk presents several complexities (USEPA, 2003b). For example, chemical interactions, which 

can lead to synergism or antagonism, are poorly understood. Also, there is a statistical 
Page 20 Examination of EPA Risk Assessment Principles and Practices 

complexity because the risks themselves are expressed as upper bound estimates of risk and not 

most likely estimates. Cogliano (1997) showed that adding upper bound estimates may make the 

resulting risk sums more conservative, but that the actual resulting risk values are not misleading 

and are probably within a factor of 2 or 3 of the estimates that would result from calculating the 

95% UCL of the sum of the most likely estimates. Cogliano states: 

(A)s the number of carcinogens increases, the sum of upper bounds becomes increasingly improbable as an 

estimate of overall risk. At the same time, however, the analysis shows that the sum of upper bounds is not 

a misleading estima te of overall risk. Obtaining similar results for different case studies sug gests that these 

conclusions a pply to mo re typical mixes of carcino gens. 

Cen tral estima tes of the o verall risk can d iffer from the sum of upp er bo unds by a fac tor of 2 - 5, as the ra tio 

between o verall uppe r and lowe r bound s decrease s ... 

In conclusion, this analysis shows that sums of plausible upper bound risk estimates do provide useful 

information about the overall risk from several carcinogens. The overall risk depends on the independence, 

additivity, and numb er of risk estimates, as well as shapes o f the underlying risk distributions. 

In response to both uncertainty and variability, EPA develops risk estimates using default 

assumptions based on empirical evidence or based on scientifically sound extrapolations. 

Further, EPA risk assessments are in fact a combination of both high-end and central tendency 

estimates. Consequently, the resulting risk estimates are expected to be on the high end of the 

range of risks but within the range of plausible outcomes. The combination of default 

assumptions is therefore reasonable, especially for independent factors, and does not result in 

exaggerated estimates. On balance, while the resulting estimates are likely to be reasonable, 

without a detailed uncertainty analysis it is not possible to determine where on the range of 

plausible outcomes the estimates actually reside. 

2.2	 General Risk Assessment Approaches Used by EPA for Public and Environmental 

Health Protection (“Public Health Protection”) 

2.2.1	 How Comprehensive Are EPA’s Risk Assessments? 

EPA cannot perform a time- and resource-intensive risk assessment for every situation 

and EPA decision. Consequently, for each risk assessment, EPA selects an approach that is 

consistent with the nature and scope of the decision being made. The appropriate approach 

depends on the needs of the decision maker and/or the role that risk information plays in the 

decision, balancing uncertainty and resources. Even using the best models and data, uncertainty 

is still inherent in the process. Given that uncertainty is inherent, there is a continuing tension 

between improving our understanding in order to make a decision and the reality of limited 

resources to perform the analysis and the desire for timely decision making. Figure 2-1 

illustrates this risk assessment continuum and the balance of resources and uncertainty as the 

assessment becomes more complex. 
Examination of EPA Risk Assessment Principles and Practices Page 21 

Figure 2-1. The Risk Assessment Continuum 

The following graphic illustrates that risk assessment can be performed with low levels of 

data and relatively little effort to develop conservative estimates of risk. Depending on the 

outcome and the needs of the risk manager, higher levels of analysis may be performed. Note 

that as one moves up the risk assessment continuum, the data needs and costs also rise. 

However, the quality of the result should also rise as well. (The following graphic is intended to 

be illustrative of the concept of tiered approaches. The actual modifications to the risk 

assessment that occur as the assessment is refined may vary from the sequence described here 

and are dependent on study-specific circumstances. In addition, a tiered approach is not 

prescriptive: one may begin with a screening-level assessment or begin with a higher level of 

analysis, as needed.) 
Page 22 Examination of EPA Risk Assessment Principles and Practices 

2.2.2 Whom Is EPA Trying To Protect? 

EPA typically cannot protect every individual but rather attempts to protect individuals 

who represent high-end exposures (typically around the 90th percentile and above) or those who 

have some underlying biological sensitivity; in doing so, EPA protects the rest of the population 

as well. In general, EPA tries to protect sensitive individuals based on normal distribution of 

sensitivities. EPA considers the most sensitive individuals where there are data, but does not 

necessarily attempt to protect “hypersensitive” individuals. The degree to which sensitive 

individuals are protected, or explicitly defined, may vary between programs based on factors 

such as the need to balance risk reductions and costs as directed and constrained by statutory 

authority. Programs may approach the problem semi-quantitatively (e.g., selecting individual 

parameter values at specified percentiles of a distribution) or qualitatively (e.g., making 

conservative assumptions to ensure protection for most individuals), though no overall degree of 

protection can be explicitly stated. 

2.2.3 Are Risk Assessment and Risk Management Separate? 

At EPA, risk assessment (evaluation of the science) and risk management (decision 

making, setting of policy) are not necessarily separate. We believe it is appropriate to involve 

decision makers from the beginning, as they typically initiate requests for risk assessments or 

analyses. Consequently, separating them entirely from the risk assessment process is neither 

logical nor desirable. Also, risk assessments typically are coordinated with the evaluations of 

economics, feasibility of remedies, and community concerns, for example, so that their results 

can be factored into decisions. EPA’s Risk Characterization Handbook (USEPA, 2000a) 

describes in detail the roles of the risk assessor and risk manager in the risk assessment process. 

Further, the NRC report on understanding risk supports the concept that risk assessment is 

conducted for the purpose of supporting risk management, and risk management considerations 

shape what is addressed in a risk characterization (NRC, 1996). 

Briefly, risk assessors are best qualified to understand the quality and nature of the data 

and to use that data to determine what the risk is, who/what is affected, the level of comfort with 

the conclusions, the uncertainty and variability inherent in the assessment, and the strengths and 

weaknesses of the assessment. In other words, they are the best qualified to make the scientific 

judgments necessary in risk assessments, including selecting models and data and assigning 

defaults where data gaps exist. Risk managers are responsible for valuing the risk and determine 

the amount of protection (as well as conservatism) to be applied in a decision. As a decision 

maker, the risk manager integrates the risk assessment with other considerations in order to make 

and justify regulatory decisions. 

However, the comments point out that many of the default assumptions and policy 

choices inherent in the risk assessment may frequently not be apparent to the risk managers. It is 
Examination of EPA Risk Assessment Principles and Practices Page 23 

the role of the risk assessors to transparently characterize such details (e.g., default assumptions, 

data selected, policy choices) so as to make clear the range of plausible risk. The risk managers, 

in their role, should inquire about the use of defaults and the choices made, if not characterized 

clearly, in order to be fully informed before using the risk assessment in making a decision. 

2.2.4 How Do “Planning and Scoping” Help Environmental Risk Assessment? 

In environmental risk assessments, planning and scoping are performed prior to the main 

analytic work. This initial work defines the questions and issues to be addressed, the analysis 

needed to address these questions/issues, and the knowledge and information needs of the 

analysis. 

Several general planning and scoping steps are relevant for many environmental risk 

assessments (see below), but specific circumstances are also important. EPA’s recently 

completed Framework for Cumulative Risk Assessment (USEPA, 2003b) provides some general 

planning and scoping steps: 

a) Defining the Purpose of the Assessment 

b) Defining the Scope of Analysis and Products Needed 

c) Agreeing on Participants, Roles and Responsibilities 

d) Agreeing on the Depth of the Assessment and the Analytical Approach 

e) Agreement on the Resources Available and Schedule 

f) Problem Formulation 

g) Developing the Conceptual Model 

h) Constructing the Analysis Plan 

Problem formulation is a systematic planning step, linked to the regulatory and policy 

context of the assessment, that identifies the major factors to be considered in a particular 

environmental assessment. The problem formulation process results in a conceptual model that 

identifies the sources, stressors, exposed populations, and the relationships among them. 

The conceptual model and the associated narrative show the basic rationale for the 

decisions made concerning the course of action for the risk assessment. Specifically, the 

conceptual model and associated narrative provide: (1) the scientific rationale for selecting the 
Page 24 Examination of EPA Risk Assessment Principles and Practices 

stressors, sources, receptors, exposed populations, exposure or environmental pathways, and 

endpoints/effects; (2) the scientific, technical, economic, or sociologic basis for the conceptual 

model; and (3) the scientific implications of additional data gathering. 

The final stage in the planning and scoping process is the development of the analysis 

plan (see discussion in USEPA, 1998a). This plan describes how hypotheses about the 

relationships among the sources, stressors, exposure conditions, populations, and adverse 

effects/endpoints presented in the conceptual model and narrative will be considered during the 

risk analysis phase of the assessment. The plan includes a rationale for which relationships are to 

be addressed and which methods and models will be used, and discusses data gaps and 

uncertainties. The plan may also compare the level of confidence needed for the management 

decision with the confidence levels expected from alternative analyses in order to determine data 

needs and evaluate which analytical approach is best. 

As stated in the EPA Science Policy Council’s 2002 Lessons Learned on Planning and 

Scoping for Environmental Risk Assessments (USEPA, 2002b), formal planning and dialogue can 

improve the final risk assessment product by making it more specific to the needs of decision 

makers and other stakeholders. Many questions and issues may be candidates for consideration 

and analysis in a risk assessment. Planning and scoping help define the boundaries of the 

analytic work (i.e. “what’s in and what’s out” of the assessment), reducing ambiguities about 

what can and cannot be done with assessment results, and helping to reduce analytic work to 

manageable segments. 

As important as how planning and scoping is done is that it is done. In general, planning 

and scoping are performed, to some degree, for all assessments. However, they may not be 

inclusive or explicit on all factors. For example, simplified analytical approaches have been 

developed and implemented over time. For each of these methods or analytical approaches, 

choices of what’s in and what’s out have already been made. These choices made may have been 

explicit when these approaches were developed, but may have become accepted over time and 

therefore no longer acknowledged. This may be especially true for the toxicity reference values. 

Some of the criticisms related to these components may stem from the lack of explicit 

characterizations of what they represent. 

2.2.5 How Does EPA Use a Screening Risk Assessment? 

Where data are sparse and uncertainty great, EPA carries out a screening risk assessment 

that tends to use default assumptions to avoid underestimating risk. These screening assessments 

typically provide high-end and bounding estimates. Pathways of trivial importance are then 

eliminated, and the remaining estimates are refined. This approach either demonstrates with 

minimal effort that no risk is large enough to consider reducing or, if that is not the case, it 

eliminates further work on refining estimates for pathways or chemicals that are clearly not 
Examination of EPA Risk Assessment Principles and Practices Page 25 

important. This is consistent with the NRC’s recommendation that EPA should use bounding 

estimates for screening assessments to determine whether further analysis is necessary (NRC, 

1994). If risks are not of concern even with these high-end or bounding screening estimates, then 

one can be fairly confident that the risks are not of concern. However, if these screening tests 

show that the risks are potentially of concern, then a more refined risk assessment may be 

warranted that uses more detailed data, models, etc., though at greater expense. An example is 

provided in section 2.2.7. These high-end screening assessments usually contain many default 

assumptions since data are generally not available. However, when usable data are available, 

they are considered instead of the defaults. 

Many comments that focused on the screening assessments of risks may have 

misunderstood the purpose of these assessments. For example, the degree to which default 

assumptions are used in exposure assessments depends on the purpose of the assessment. By 

their very nature, screening assessments are broad in scope and based on relatively sparse data. 

In these cases, EPA attempts to clearly identify that the assessments are for screening purposes 

and to explain the meaning of the results and the utility of the assessment in the context of 

whatever decision is at hand. 

2.2.6	 What Happens if EPA Identifies a Potential Risk That Needs To Be 

Addressed After a Screening Risk Assessment? 

When a screening assessment identifies the potential for a non-trivial risk, EPA decides if 

pursuing that risk is appropriate based on its current priorities and available resources. If the 

Agency decides to pursue the risk, more detailed, refined risk assessments are then performed, 

though the degree of refinement (i.e., where the risk assessment falls along the continuum shown 

in figure 2-1) depends on the type of decision, the available resources, and the needs of the 

decision maker. For those pathways or chemicals that were shown to be non-trivial by bounding 

estimates, we work to refine our estimates of exposure and dose. At this point we estimate 

exposures, doses, and responses that fall on the distribution of actual exposures pertinent to the 

population under study. In performing this continued analysis, we use a combination of data, 

ranges of data, distributions of data, and assumptions about each of the factors needed to estimate 

risk. Generally, we perform both central tendency and high-end estimates (and, increasingly, we 

develop fully probabilistic risk distributions). Each of these estimates is surrounded by 

uncertainty (perhaps unquantifiable); the degree of uncertainty depends on the quality and 

comprehensiveness of the available data. 

For the more refined assessment, EPA would like to use appropriate and available data to 

generate a more data-based assessment. This poses great difficulty when the data are not 

available and/or adequate. If data are simply not available, then we usually employ basic default 

assumptions. More frequently, data are available but deemed inadequate by EPA. This creates a 

potentially confrontational situation if the generator of the data claims the data are adequate. 
Page 26 Examination of EPA Risk Assessment Principles and Practices 

Peer review has been extremely helpful for determining the adequacy of the data, and EPA 

makes extensive use of this process. When peer review judges the data to be inadequate, we 

generally fall back on using defaults to help assess the risk. This situation probably is the basis 

for many of the comments that EPA does not use current data and/or chemical- or site-specific 

data. In fact, we do try to use the most relevant information as validated by peer review. 

When exposure or dose estimates have sufficiently narrow uncertainty relative to the 

needs of the decision maker, we can develop the final risk assessment. Otherwise, the data or 

assumptions used usually have to be even further refined, if resources allow, in an attempt to 

further reduce uncertainty and bring the estimated exposure or dose closer to the actual values in 

the population. Refining the estimates usually requires that new data be considered. These data 

may come from other studies in the literature, information previously developed for a related 

purpose and adapted, or new survey, laboratory, or field data. The decision about which 

particular parts of the information base to refine should be based both on which data will most 

significantly reduce the uncertainty of the overall exposure or dose estimate of interest, and on 

which data are in fact obtainable either technologically or within resource constraints. After 

refinement of the estimate, we again determine whether the estimates provided will be sufficient 

to answer the questions posed to an acceptable degree, given the uncertainties that may be 

associated with those estimates. Refinements proceed iteratively until the assessment provides 

an adequate answer for the decision maker within the resources available. 

2.2.7 How Are High-End Exposures Reflected in EPA Evaluations? 

Although populations experience a range (or distribution) of exposures, a number of 

environmental statutes require that EPA consider those exposures at the high end of the 

distribution when making certain decisions. Utilizing high end exposures as one component of 

the risk decision making process helps to ensure equitable protection across an exposed 

population (i.e., protecting a high end person in a population helps ensure protection of most of 

the population). 

One example of this is EPA’s approach for evaluating a population’s high end exposure 

to hazardous air pollutants (HAPs). This has sometimes been referred to as evaluation of “the 

porch potato” (i.e., the assumption that someone lives outdoors at the point of maximum 

concentration at or beyond the fence line of a facility for 24 hours a day for a lifetime). This 

section discusses the statutory and analytical frameworks that provide the basis for EPA’s risk 

assessments for HAPs pursuant to certain provisions of the Clean Air Act (CAA). 
Examination of EPA Risk Assessment Principles and Practices Page 27 

What Is the Statutory Basis for EPA’s Selection of Exposure Scenarios When Assessing 

Hazardous Air Pollutant Risks? 

The Clean Air Act identifies the risk to the individual most exposed (IME) as the risk of 

interest when making certain decisions about the regulation of HAPs. Specifically, the 1990 

CAA Amendments direct EPA to consider risk to the IME when determining whether a source 

category may be deleted from the list of sources of HAPs (Section 112(c)(9)(B)(i)), and when 

determining whether residual risk standards are necessary (Section 112(f)(2)(A)). Residual risk 

standards are the mechanism that Congress provided EPA for addressing public health and 

environmental risks that may remain after the regulated community has implemented technologybased standards for the control of HAPs. 

In addition, Section 112(f)(2)(B) of the 1990 CAA Amendments incorporates by 

reference the use of a two-step risk assessment framework for setting residual risk standards 

under Section 112(f)(2)(A). (Note: EPA has not yet proposed any residual risk standards, but is 

working on a number of residual risk determinations and expects to issue the first proposal in 

2004.) The two-step framework was articulated in EPA’s 1989 Benzene NESHAP (54 Federal 

Register 38044) and consists of: 

1)	 A first step, in which EPA ensures that risks are “acceptable.” As explained in the 

Benzene NESHAP, in this step EPA generally limits the maximum individual risk, 

or “MIR,” to no higher than approximately 1 in 10 thousand. The benzene 

NESHAP defines the MIR as the estimated risk that a person living near a plant 

would have if he or she were continuously exposed to the maximum pollutant 

concentrations for a lifetime (70 years). 

2)	 A second step, in which EPA establishes an “ample margin of safety.” In this 

step, EPA strives to protect the greatest number of persons possible to an 

estimated individual excess lifetime cancer risk level no higher than 

approximately 1 in 1 million. 

In judging whether risks are acceptable and whether an ample margin of safety is 

provided, the benzene NESHAP states that EPA will consider not only the magnitude of 

individual risk, but “the distribution of risks in the exposed population, incidence, the science 

policy assumptions and uncertainties associated with the risk measures, and the weight of 

evidence that a pollutant is harmful to health.” Therefore, decisions under Section 112(f)(2)(A) 

typically will include consideration of both population risk and individual risk, as well as other 

factors. (Note that the ample margin of safety analysis in the second step of the residual risk 

framework also includes consideration of additional factors relating to the appropriate level of 

control, “including costs and economic impacts of controls, technological feasibility, 

uncertainties, and any other relevant factors.”) 
Page 28 Examination of EPA Risk Assessment Principles and Practices 

What Analytical Framework Does EPA Use to Estimate the “IME” and“MIR?” 

Risk assessment for air toxics involves using models to estimate HAP concentrations in 

air (and other media, as necessary), and then combining these HAP concentrations with other 

exposure assumptions and measures of pollutant toxicity to estimate risk (usually 

deterministically). As a result, OAR’s risk estimates are ultimately a function of the values 

selected for numerous parameters, including: HAP-specific emission factors, meteorological 

parameters (e.g., wind speed, wind direction, precipitation, temperature), stack parameters (e.g., 

height, diameter, temperature, exit velocity), distance to receptor, and duration of exposure (e.g., 

residential occupancy period). 

The exposure scenario that EPA uses to estimate the MIR is provided in the Benzene 

NESHAP. The Benzene NESHAP specifies that the MIR be based on exposure to the maximum 

pollutant concentrations for 70 years. The Benzene NESHAP indicates that it is appropriate to 

account for habitability when identifying the location of maximum pollutant concentrations. 

Therefore, EPA may characterize the maximum off-site annual average concentration in 

habitable areas (e.g., excluding such areas as lakes) to estimate the MIR. 

OAR typically uses a tiered analytical approach when estimating risk to the IME. As 

discussed above, because there are many parameters that influence an individual’s exposure, it is 

not possible to identify the actual individual in the population who is most exposed. OAR 

recognizes, however, that it is very unlikely that there exists an individual who should be 

characterized using the worst-case values of all exposure and toxicity parameters. Nevertheless, 

a tiered analysis often begins with a “worst-case” or bounding analysis that generally sets 

parameters at values that maximize the estimate of risk (e.g., exposure is assumed to continue for 

a lifetime). If risks estimated using such an analysis are not of concern, then there is no need to 

refine the analysis further, and EPA may proceed with the appropriate action (i.e., EPA may 

delete the source pursuant to 112(c)(9)(B)(i) or make a determination not to propose residual risk 

standards under 112(f)(2)(A)). If risks estimated in the initial analysis are of potential concern, 

analysts may make successive refinements in modeling methodologies and input data to derive 

successively less conservative, more realistic, estimates of the risk to the IME. 

In refining the risk estimate for the IME, the exposure assumptions for which OAR has 

the best alternative information generally are modified first. For example, where facility-specific 

information is available, OAR will use the actual locations of residences (e.g., placing the 

receptor in the geographic center of a populated census block near a facility) to estimate the 

concentrations of HAPs in air. OAR also may conduct additional data gathering and analysis to 

further refine the characteristics of the actual emissions and emission points. The goal of 

successive tiers of the analysis is to start with a bounding or worst-case exposure estimate and 

then move away from the worst-case parameters until the combination of parameter values 

represents, in the judgment of the assessor, the exposure experienced by the individual in the 
Examination of EPA Risk Assessment Principles and Practices Page 29 

population who is most exposed. However, where the risk estimates are below the level of 

concern, there is no need to conduct additional analysis to refine the estimates further, even if the 

estimates might still overestimate the risk to the individual in the population who is most 

exposed. 

As described above, the risk estimate is a function of the values of a number of 

parameters. In refining the bounding or worst-case risk analysis, OAR has not modified the 

assumption of 70-year, 24-hour per day, outdoor exposure. Although OAR recognizes that the 

majority of people do not reside outdoors and in one location for their entire lives, and that there 

is a general trend of increased population mobility, OAR believes that the data available for 

refining assumptions for exposure duration and frequency are less certain than the data available 

for refining other parameter values. For example: 

a) Residential occupancy periods may be influenced by factors (e.g., economic, 

geographic) that may cause local population mobility patterns to differ from 

national estimates of population mobility. 

b) If a source of concern occurs in the majority of communities in the country, then it 

is possible that an individual may be exposed to the source for a longer period of 

time than one might predict using national estimates of population mobility. That 

is, even though an individual moves, the individual’s new residence may be 

located near a similar source of concern. 

c) If a single source impacts a large geographic area, then it is possible that an 

individual may move or travel from one point of exposure to the source to another 

point of exposure to the same source. 

d) Exposure to HAPs may not diminish when individuals are indoors. Empirical 

data for many pollutants show that long-term average indoor concentrations of 

outdoor air pollutants are roughly equivalent to long-term average outdoor 

concentrations of those pollutants (e.g., Sexton et al., 2004). 

Moreover, OAR is cognizant that NRC acknowledged some of these issues in its 1994 report, 

which recommended that: “EPA should use the mean of current life expectancy as the 

assumption for the duration of individual residence time in a high-exposure area, or a distribution 

of residence times which accounts for the likelihood that changing residences might not result in 

significantly lower exposure. Similarly, EPA should use a conservative estimate for the number 

of hours a day an individual is exposed, or develop a distribution of the number of hours per day 

an individual spends in different exposure situations.” 
Page 30 Examination of EPA Risk Assessment Principles and Practices 

3. UNCERTAINTY AND VARIABILITY 

3.1 Overview 

Uncertainty and variability exist in all risk assessments. Even at its best, risk assessment 

does not estimate risk with absolute certainty. Thus, it is important that the risk assessment 

process handle uncertainties in a predictable way that is scientifically defensible, consistent with 

the Agency’s statutory mission, and responsive to the needs of decision makers (NRC, 1994). 

Instead of explicitly quantifying how much confidence there is in a risk estimate, EPA attempts 

to increase the confidence that risk is not underestimated by using several options to deal with 

uncertainty and variability when data are missing. For example, in exposure assessment, the 

practice at EPA is to collect new data, narrow the scope of the assessment, use default 

assumptions, use models to estimate missing values, use surrogate data (e.g., data on a parameter 

that come from a different region of the country than the region being assessed), and/or use 

professional judgment. The use of individual assumptions can range from qualitative (e.g., 

assuming one is tied to the residence location and does not move through time or space) to more 

quantitative (e.g., using the 95th percentile of a sample distribution for an ingestion rate). This 

approach can also fit the practice of hazard assessment when data are missing. Confidence in 

ensuring that risk is not underestimated has often been qualitatively ensured through the use of 

default assumptions. 

EPA has been increasingly concerned about characterizing uncertainty in its risk 

estimates. EPA’s 1986 set of Risk Assessment Guidelines explicitly stated the importance of 

characterizing uncertainty. EPA’s Exposure Assessment Guidelines developed this theme further 

for the exposure assessment part of risk assessment. EPA’s Risk Characterization Policy 

provided even more direction for describing uncertainty in risk estimates. For probabilistic 

analysis specifically, EPA made significant efforts in recent years to use probabilistic techniques 

to characterize uncertainty; these include the March 1997 Guiding Principles for Monte Carlo 

Analysis (USEPA, 1997b), the May 1997 Policy Statement (USEPA, 1997c), and the December 

2001 Superfund document Risk Assessment Guidance for Superfund: Volume III — Part A, 

Process for Conducting Probabilistic Risk Assessment (USEPA, 2001a) 

3.2 Uncertainty and Variability 

3.2.1 What Is Uncertainty? 

Uncertainty can be defined as a lack of precise knowledge as to what the truth is, whether 

qualitative or quantitative. 
Examination of EPA Risk Assessment Principles and Practices Page 31 

Numerous schemes for classifying uncertainty have been proposed. The preferred 

approach of the NRC (1994) focused on two broad categories: parameter uncertainty and model 

uncertainty. These are defined below in sections 3.2.2 and 3.2.3. 

3.2.2	 What Is Parameter Uncertainty? 

Risk assessments depict reality interpreted through mathematical representations that 

describe major processes and relationships. Process or mechanistic models use equations to 

describe the processes that an environmental agent undergoes in the environment in traveling 

from the source to the target organism. Mechanistic models have also been developed to 

represent the toxicokinetic and toxicodynamic processes that take place inside the organism, 

leading to the toxic endpoint. The specific parameters of the equations found in these models are 

factors that influence the release, transport, and transformation of the environmental agent, the 

exposure of the target organism to the agent, transport and metabolism of the agent in the body, 

and interactions on the cellular and molecule levels. Empirical models are also used to define 

relationships between two values, such as the dose and the response. Uncertainty in parameter 

estimates stem from a variety of sources, including: 

a)	 Measurement errors: 

1)	 Random errors in analytical devices (e.g., imprecision of continuous 

monitors that measure stack emissions). 

2)	 Systemic bias (e.g., estimating inhalation from indoor ambient air without 

considering the effect of volatilization of contaminants from hot water 

during showers). 

b)	 Use of surrogate data for a parameter instead of direct analysis of it (e.g., use of 

standard emission factors for industrialized processes). 

c)	 Misclassification (e.g., incorrect assignment of exposures of subjects in historical 

epidemiologic studies due to faulty or ambiguous information). 

d)	 Random sampling error (e.g., estimation of risk to laboratory animals or exposed 

workers in a small sample). 

e)	 Non-representativeness (e.g., developing emission factors for dry cleaners based 

on a sample of “dirty” plants). 
Page 32 Examination of EPA Risk Assessment Principles and Practices 

3.2.3	 What Is Model Uncertainty? 

Model uncertainties arise because of gaps in the scientific theory that is required to make 

predictions on the basis of causal inferences. Common types of model uncertainties in various 

risk assessment–related activities include: 

a)	 Relationship errors (e.g., incorrectly inferring the basis of correlations between 

chemical structure and biological activity). 

b)	 Oversimplified representations of reality (e.g., representing a three-dimensional 

aquifer with a two-dimensional mathematical model). 

c)	 Incompleteness, i.e., exclusion of one or more relevant variables (e.g., relating 

asbestos to lung cancer without considering the effect of smoking on both those 

exposed to asbestos and those unexposed). 

d)	 Use of surrogate variables for ones that cannot be measured (e.g., using wind 

speed at the nearest airport as a proxy for wind speed at the facility site). 

e)	 Failure to account for correlations that cause seemingly unrelated events to occur 

more frequently than expected by chance (e.g., two separate components of a 

nuclear plant are both missing a particular washer because the same newly hired 

assembler put them together). 

f)	 Extent of (dis)aggregation used in the model (e.g., whether to break up fat 

compartment into subcutaneous and abdominal fat in a physiologically based 

pharmacokinetic, or PBPK, model). 

Model uncertainty is often difficult to quantify. Further, it is inherent in risk assessment 

that seeks to capture the complex processes impacting release, environmental fate and transport, 

exposure, and exposure-response. EPA’s models are often incomplete and knowledge of specific 

processes limited. As a result, EPA relies on specific default assumptions as a response to 

uncertainty. Again, these represent scientifically plausible choices that are intended to more 

likely guard against underestimating risks. 

3.2.4	 How Does Variability Differ From Uncertainty? 

Variability is considered with uncertainty since, like uncertainty, it does not allow for a 

specific single correct value. However, they differ significantly in terms of their impact and role 

in risk assessments. Uncertainty — the lack of knowledge — can be reduced through additional 
Examination of EPA Risk Assessment Principles and Practices Page 33 

investigation. Variability is inherent heterogeneity across space, in time, or among individuals; it 

cannot be reduced with additional investigation, only better understood or characterized. 

“Human variability” refers to person-to-person differences in biological susceptibility or 

in exposure. Although both human variability and uncertainty can be characterized as ranges or 

distributions, they are fundamentally different concepts. Again, uncertainty can be reduced by 

further research that supports a model or improves a parameter estimate, but human variation is a 

reality that can be better characterized, but not reduced, by further research. (Note that fields 

other than risk assessment use “variation” or “variability” to mean dispersion about a central 

value, including measurement errors and other random errors that risk assessors address as 

uncertainty.) 

EPA addresses variability by assessing the risk to the sensitive portions of the population 

or ecosystem (see section 3.6 for fuller discussion). Accordingly, EPA typically makes explicit 

choices to characterize the risks at the upper end of the expected distribution. Typically, EPA 

focuses on the critical parameter or assumption in any particular aspect of an assessment and 

based on the particular needs of the decision maker, including the mandates and constraints in 

statutory language, and selects the appropriate choice — such as the maximum or a specified 

percentile. In general, EPA’s approach to variability has focused on exposure (e.g., 

characterizing the risks to the individual most exposed within the Air Toxics or Hazardous Air 

Pollutant program, section 2.2.7), but the approach may also be reflected in the toxicity part of 

the assessment (e.g., UF for human variability). 

3.3 Characterizing Uncertainty and Variability 

One of the major comments on EPA risk assessment practices is that they do not 

characterize uncertainty and variability transparently enough. This is an issue EPA is attempting 

to address. The comments also encourage EPA to begin using probabilistic analysis to describe 

quantitatively the uncertainty in risk estimates where appropriate. 

3.3.1 Why Is Characterizing Uncertainty and Variability Important? 

The very heart of risk assessment is the responsibility to use whatever information is at 

hand or can be generated to produce an estimate, a range, a probability distribution — whatever 

best expresses the present state of knowledge about the effects of some hazard in some specific 

setting. To ignore the uncertainty in any process is almost sure to leave critical parts of the 

process incompletely examined and hence to increase the probability of generating a risk estimate 

that is incorrect, incomplete, or misleading (NRC, 1994). 

The NRC (1994) further noted that risk assessments that do not pay sufficient attention to 

uncertainty are vulnerable to four common, potentially serious pitfalls: 
Page 34 Examination of EPA Risk Assessment Principles and Practices 

a) They do not allow for optimal weighting of the probabilities and consequences of 

error. 

b) They do not permit a reliable comparison of alternative decisions. 

c) They fail to communicate the range of control options that would be comparable 

with different assessments of the true state of nature. 

d) They preclude the opportunity for identifying research initiatives. 

Further, uncertainty analysis will play a more prominent and formal role in regulatory 

decision making. For example, OMB’s recent revisions to its regulatory analysis guidelines 

(USOMB, 2003c) state that formal quantitative uncertainty analysis be performed for economic 

assessment in support of overall regulatory analysis. Whenever possible, appropriate statistical 

techniques should be used to determine the probability distribution of relevant outcomes. For 

major rules involving annual economic effects of $1 billion or more, a formal quantitative 

analysis of uncertainty is required. The OMB guidelines outline analytical approaches of varying 

levels of complexity, which could be used for uncertainty analysis such as qualitative disclosure, 

numerical sensitivity analysis, and formal probabilistic analysis (required for rules with impacts 

greater than $1 billion). 

3.3.2 When Should EPA Conduct Uncertainty Analysis? 

As the NRC notes, EPA’s analysis of uncertainty has tended to be piecemeal and highly 

focused on an assessment’s sensitivity to the accuracy of a few specified assumptions rather than 

a full exploration of the process from data collection to final risk assessments (NRC, 1994). 

Further, EPA tends to conduct uncertainty analysis a posteriori, focusing on specific assumptions 

or variables. The level of treatment again depends on the decision and the value such an analysis 

would provide to that decision. EPA believes this is an appropriate approach to balance 

resources with value added. Some have suggested that EPA should integrate analysis and 

uncertainty into the basic development of the primary study (see USEPA, 2000e). Until such 

methods and supporting data are developed, though, it is not feasible to do a full and integrated 

assessment for every analysis. Further, in most instances EPA is not the data developer: much 

EPA analysis is based upon third-party literature. 

Ideally, an uncertainty analysis would be built into the study design; hopefully, as 

methods and supporting data continue to be developed, we will move toward such an approach in 

the future. 
Examination of EPA Risk Assessment Principles and Practices Page 35 

3.3.3	 What Is an Appropriate Level of Uncertainty Analysis? 

Over the years, improved computer capabilities have created more opportunities to 

characterize uncertainty. As a result, advocates promote such characterization in all cases. We 

need to be judicious in which methods we apply, such as Monte Carlo analysis. Uncertainty 

analysis is not a panacea, and full formal assessments can still be time- and resource-intensive. 

Further, the time and resources needed to collect an adequate database for such analyses can be a 

problem. While uncertainty analysis arguably provides significant information to aid in decision 

making, its relative value is case-specific and depends on the characteristics of the assessment 

and the decision being made. In some cases, a full probabilistic assessment may add little value 

relative to simpler forms. This may occur where more detailed uncertainty analysis (or analysis 

focused on non-critical uncertainties) does not provide information which has any impact on the 

overall decision. 

Accordingly, EPA’s practice is to use a “tiered approach” to conducting uncertainty 

analysis; that is, EPA starts as simply as possible (e.g., with qualitative description) and 

sequentially employs more sophisticated analyses (e.g., sensitivity analysis to full probabilistic), 

but only as warranted by the value added to the analysis and the decision process. Questions 

regarding the appropriate way to characterize uncertainty include: 

a)	 Will the quantitative analysis improve the risk assessment? 

b)	 What are the major sources of uncertainty? 

c)	 Are there time and resources for a complex analysis? 

d)	 Does this project warrant this level of effort? 

e)	 Will a quantitative estimate of uncertainty improve the decision? How will the 

uncertainty analysis affect the regulatory decision? 

f)	 How available are the skills and experience needed to perform the analysis? 

g)	 Have the weaknesses and strengths of the methods involved been evaluated? 

h)	 How will the uncertainty analysis be communicated to the public and decision 

makers? 
Page 36 Examination of EPA Risk Assessment Principles and Practices 

3.3.4 With What Precision Should EPA Results Be Reported? 

Uncertainty is inherent in almost all the data EPA reviews. However, at the end of many 

risk assessments, values are provided that appear to have several significant digits. This implies 

a certain level of precision that probably is not justified. EPA is currently addressing this 

practice of characterizing the values. 

With the increased use of computational methods and readily available software, 

estimates can be reported with apparently unlimited precision. Nonetheless, the precision to 

which a value is reported should be commensurate with the level of confidence in that number. 

EPA has been criticized in the past for reporting risk estimates with unnecessary and 

unsupportable degrees of precision. Often, this is the result of automated reporting within 

commercial software packages, or of incremental stages of analysis that are not reviewed and 

edited in accordance with their incremental nature. 

EPA agrees that results should be reported to their proper level of precision and the 

value(s) well characterized. EPA guidance on presentation of risk information comes from 

various documents developed since 1989. The Risk Assessment Guidance for Superfund: 

Volume I — Part A (USEPA, 1989a) provides guidance on the number of significant figures to 

use in presenting cancer risks in the risk characterization. RAGS Part A’s page 8-7 (Exhibit 8-2, 

footnote b) states that “All cancer risks should be expressed as one significant figure only.” In 

addition, the Risk Assessment Guidance for Superfund: Volume 1 — Part D (USEPA, 2001b) 

provides standardized tables for presenting risk information. The Record of Decision guidance 

provides standardized tables for presenting risk information (USEPA, 1999a); the same 

document provides guidance on presenting non-cancer hazard indices. 

In a final risk description, all cancer risks and non-cancer hazard indices should be 

presented with one significant figure only. More figures, however, should be carried along the 

way to minimize rounding errors and to make it possible for others to verify calculations. The 

number of significant figures presented during a risk assessment changes to reflect the needs of 

the assessment. For example, risk information presented in RAGS Part D tables (tables 6 

through 9 in RAGS) typically present cancer risk information with two significant figures. This 

summary information allows risk assessors reviewing the document to check the accuracy of the 

cancer risk calculations and provide appropriate comments. 

3.4 Issues in Characterizing Uncertainty and Variability 

EPA has been examining its practice of uncertainty analysis over the years. Generally 

speaking, EPA has applied a qualitative approach to characterizing uncertainty (and variability 

for that matter). Quantitative characterization has been used more unevenly. Also, EPA 
Examination of EPA Risk Assessment Principles and Practices Page 37 

typically uses deterministic approaches to characterize risk — although, increasingly often, EPA 

applies probabilistic techniques for characterization of risk, usually within exposure assessments. 

3.4.1	 What Is the Issue of Deterministic Versus Probabilistic Approaches? 

Risk assessments may consider both non-cancer and cancer endpoints when data are 

available. A non-cancer assessment includes an oral reference dose (RfD) and an inhalation 

reference concentration (RfC). A cancer assessment includes a cancer weight-of-evidence 

narrative statement and, if data permit, a quantitative cancer risk estimate. Many of the risk 

estimates from these assessments are found in the Agency’s Integrated Risk Information System 

(IRIS) as well as in EPA offices that also develop health assessments.. 

As an example for the practice of characterizing uncertainty, we examine the traditional 

approach to dose-response assessment analysis of non-cancer endpoints (RfD or RfC). The 

approach is to identify the no-observed-adverse-effect level (NOAEL) and lowest-observedadverse-effect level (LOAEL) from an appropriate study. The NOAEL (or LOAEL, if a NOAEL 

is not available) is adjusted downward by uncertainty factors (UFs) intended to account for 

uncertainties in the available data, producing an exposure that is likely to pose no increased risk 

of adverse effects for chronic exposure. EPA currently also uses the benchmark dose (BMD) 

approach to deriving a RfD/RfC when appropriate data are available. 

While there may be significant statistical analysis and consideration of (both qualitative 

and quantitative) uncertainty, an estimated reference value or exposure is largely developed via a 

deterministic process and yields a deterministic estimate. Part of this approach is the use of UFs 

that are applied to adjust for uncertainties in extrapolating from the type of study serving as the 

basis for the RfD/RfC to the situation of interest for the risk assessment. UFs are used to account 

for each of the extrapolations used in the assessment, such as use of animal data (interspecies 

extrapolation) and protection of susceptible individuals (intraspecies variability). 

A deterministic value for “a dose or exposure not expected to cause adverse effects” may 

convey an inappropriate degree of precision, and some think that a single number is not 

appropriate. A single number is frequently presented as “the risk assessment”; it is clear this is 

not an adequate way to characterize any potential risk. Several considerations need to be 

characterized to increase confidence in the assessment (i.e., to reduce uncertainty): 

a)	 The degree of confidence needed in stating that susceptible/sensitive individuals 

are affected or not above the resulting exposures. 

b)	 The condition of the database from which the original BMD, NOAEL, or LOAEL 

was developed (i.e., the quality of the individual study on which the value is 

based). 
Page 38 Examination of EPA Risk Assessment Principles and Practices 

c)	 The numerical values for the UFs employed in calculating the RfC/RfD. 

d)	 The procedure by which the BMD, NOAEL, or LOAEL from the experimental 

system may have been adjusted to account for known differences in the human 

population of interest (e.g., dosimetric adjustments). 

e)	 The shape of the dose-response curve. 

A probabilistic approach could be advantageous in that it would eliminate the need to 

define a single value and might be less likely to imply undue precision. In such a framework, a 

probability distribution would be used to express the belief that any particular value represents 

the dose or exposure concentration that would pose no appreciable risk of adverse effects. In 

contrast with the imprecision of a deterministic approach, several researchers still attempt to 

interpret what is meant by the RfD or RfC and what is the degree of protection afforded (e.g., 

Swartout et al., 1998; Baird et al., 1996; Gaylor and Kodell, 2000). 

3.4.2	 What Is the Importance of Quantitative Characterization of Uncertainty in 

Dose-Response? 

EPA has stressed the importance of attention to uncertainty in risk assessment. We quote 

extensively here from the successive versions of EPA’s Cancer Guidelines, since they articulate 

the issue well in terms of cancer risk assessment. Many of the points apply to this issue for risk 

assessment in general. The 1986 Cancer Guidelines state: 

It should be emphasized in every quantitative risk estimation that the results are uncertain. Uncertainties 

due to experimental and epidemiologic variability as well as uncertainty in the exposure assessment can be 

impo rtant. T here a re ma jor un certainties in extra pola ting bo th from anima ls to hum ans and from high to 

low doses. There are important species differences in uptake, metabolism, and organ distribution of 

carcinogens, as well as species and strain differences in target-site susceptibility. Human populations are 

variable with respe ct to genetic constitution, diet, occup ational and ho me enviro nment, activity patterns, 

and other cultural factors. Risk estimates should be presented together with the associated hazard 

assessment (section III.C.3.) to ensure that there is an appreciation of the weight of evidence for 

carcinoge nicity that underlies the quantitative risk estimates. 

At that time EPA recognized the need for further development of methodologies to 

address uncertainties: 

It is also recognized that there is a need for new methodology that has not been addressed in this document 

in a number of areas, e.g., the characterization of uncertainty. As this knowledge and assessment 

methodology are developed, these Guidelines will be revised whenever appropriate. 
Examination of EPA Risk Assessment Principles and Practices Page 39 

The 1999 draft revised Cancer Guidelines contain substantial additional discussion about 

uncertainty. In the key area of dose-response assessment, they state: 

The characterization presents the results of analyses of dose data, of response data, and of dose-response. 

When alternative approaches are plausible and persuasive in selecting dose data, response data, or 

extrapolation procedures, the characterization follows the alternative paths of analysis and presents the 

results. 

This discussion goes on to address specific areas where uncertainties, including 

quantitative uncertainties, are to be considered:“Uncertainty analyses, qualitative or quantitative 

if possible, are highlighted in the characterization.” The guidelines continue with a list of 

specific factors to include in the characterization. 

The 2003 draft final Cancer Guidelines contain a further expansion in the discussion of 

uncertainty issues: 

Some aspects of model uncertainty that should be addressed in an assessment include the use of animal 

models as a surrogate for humans, the influence of cross-species differences in metabolism and physiology, 

the use of effects observed at high doses as an indicator of the potential for effects at lower doses, the effect 

of using linear or nonlinear extrapolation to estimate risks, the use of small samples and subgroups to make 

inferences about entire human populations or subpopulations with differential susceptibilities, and the use of 

experimental exposure regimens to make inferences about different human exposure scenarios (NRC, 

2002). 

Toxicokinetic and toxicodynamic models are generally premised on site concordance across species ... The 

assessment should discuss the relevant data that bear on this form of model uncertainty. ... 

Probabilistic risk assessment, informed by expert judgment, has been used in exposure assessment to 

estimate human variation and uncertainty in lifetime average daily dose. Probabilistic methods can be used 

in this exposure assessment application because the pertinent variables (for example, concentration, intake 

rate, exposure duration, and body weight) have been identified, their distributions can be observed, and the 

formula for combining the variables to estimate the lifetime average daily dose is well defined (see USEPA, 

1992a). Similarly, probabilistic methods can be applied in dose-response assessment when there is an 

understanding of the important parameters and their relationships, such as identification of the key 

determinants of human variation (for example, metabolic polymorphisms, hormone levels, and cell 

replication rates), observation of the distributions of these variables, and valid models for combining these 

variables. With appropriate data and expert judgment, formal approaches to probabilistic risk assessment 

can be applied to provide insight into the overall extent and dominant sources of human variation and 

uncertainty. In doing this, it is important to note that analyses that omit or underestimate some principal 

sources of variation or uncertainty could provide a misleadingly narrow description of the true extent of 

variation and uncertainty and give decision-makers a false sense of confidence in estimates of risk. 

Specification of joint probability distributions is appropriate when variables are not independent of each 

other. In each case, the assessment should carefully consider the questions of uncertainty and human 

variation and discuss the extent to which there are data to address them. 

Probabilistic risk assessment has been used in dose-response assessment to determine and distinguish the 

degree of uncertainty and variability in toxicokinetic and toxicodynamic modeling. Although this field is 
Page 40 Examination of EPA Risk Assessment Principles and Practices 

less advanced than probabilistic exposure assessment, progress is being made and these guidelines are 

flexible enough to accom modate continuing advances in these approaches. 

3.4.3 How Does EPA Use Probabilistic Analysis? 

EPA cancer and other risk assessments have not included full probabilistic uncertainty 

analyses to date, primarily due to the need to develop relevant probability distributions in the 

toxicity part of risk assessment. However, quantitative statistical uncertainty methods are 

routinely applied in evaluation of fitting of dose-response models to tumor data, and quantitative 

uncertainty methods have been used to characterize uncertainty in pharmacokinetic and 

pharmacodynamic modeling. 

As stated in EPA’s Policy for Use of Probabilistic Analysis in Risk Assessment (1997c): 

For human health risk assessments, the application of Monte Carlo and other probabilistic techniques has 

been limited to exposure assessments in the majority of cases. The current policy and associated guiding 

princ iples ar e not inte nde d to a pply to dose -respo nse evaluations for human health risk assessm ent until this 

application of probabilistic analysis has been studied further. 

The EPA’s Science Advisory Board (SAB) suggests that the Agency reconsider this 

policy (see USEPA, 2000e) for dose-response evaluations for human health risk assessments. 

Generally, “probabilistic analysis” is a means for describing the uncertainty in risk 

estimates by characterizing the uncertainty and population variability in the individual steps by 

probability distributions. Thus, the likelihood of each risk is quantitatively characterized in the 

resulting estimates. This is generally implemented by a “Monte Carlo” approach, which 

performs a computer simulation to produce the estimates. 

EPA provided guidance on the use of probabilistic analysis and made significant efforts 

to use probabilistic techniques to characterize uncertainty. In recent years, as stated above, EPA 

developed the March 1997 Guiding Principles for Monte Carlo Analysis, the May 1997 Policy 

Statement, and the December 2001 Superfund document Risk Assessment Guidance for 

Superfund: Volume III — Part A, Process for Conducting Probabilistic Risk Assessment. Before 

performing a probabilistic risk assessment, one should carefully examine the available data to 

determine if they are adequate for the analysis and to consider the costs associated with the 

probabilistic risk assessment relative to the proposed action. 

There is still more work to be done, particularly in the hazard assessment area and in 

combining the hazard assessment and exposure assessment results. Probabilistic risk assessment 

could probably be used more frequently; it could provide useful information for many risk 

assessments beyond screening-level assessments. Probabilistic risk assessment incorporating the 

entire process, including the toxicity part of the risk assessment and not only the exposure 
Examination of EPA Risk Assessment Principles and Practices Page 41 

component, would be particularly useful. In addition, probabilistic analysis may avoid some 

potential problems of apparent overestimation of risk estimates from multiplying UFs in a 

deterministic risk assessment. (Most of the time, there is more of a tendency for the estimate to 

be higher than the “true” risk rather than lower. In individual cases, however, it is most certainly 

possible that the final, overall risk estimates will underestimate “true” risk for a particular risk 

assessment.) 

We recognize that probabilistic analysis is no panacea. It will not necessarily result in 

different outcomes or decisions, nor are its implementation and use simple processes. The 

accuracy of probabilistic analysis will still depend upon the quality of the data used for the 

analysis. Moreover, in some situations, a probabilistic analysis may not be appropriate. For 

example, at cleanup sites (such as Superfund, Resource Conservation and Recovery Act, or 

Brownfields sites), the resident populations may not have the wide-ranging susceptibilities that 

would warrant conducting a probabilistic risk assessment. 

Although EPA provided guidance for some use of probabilistic techniques, and facilitated 

its use via the May 1997 policy statement, the Task Force agrees that EPA should attempt, where 

appropriate, to make more use of probabilistic techniques: they could provide useful information 

on both variability in exposures and uncertainty in the estimates for many risk assessments that 

go beyond screening-level assessments. In addition to making variability and uncertainty more 

transparent (e.g., sometimes allowing disaggregation of the effects of variability and uncertainty 

by “nested” simulations), probabilistic analysis can provide more consistency in characterizing 

risks. Further, probabilistic analysis may avoid some potential problems of upwardly biased 

estimates of risk from combining uncertainty/variability input variables in a deterministic risk 

assessment. (By “upwardly biased,” it is meant that, on average, there is more of a tendency for 

the estimate to be higher than the “true” risk rather than lower. In individual cases, however, it is 

most certainly possible that the final, overall risk estimates will underestimate “true” risk for a 

particular risk assessment.) 

3.4.4 How Does Research Help Reduce Uncertainty? 

The NRC strongly stated (NRC, 1994) and EPA agrees that risk assessment is intended to 

be an iterative process that promotes further investigation (research) to reduce uncertainties. 

Once available, this improved information is applied to the risk assessment process to refine the 

assessment. The information obtained from research improves confidence in the risk assessment 

and reduces the associated uncertainty. Further, the use of defaults does not preclude, reduce, or 

eliminate the push for new science. 

Such an approach suggests that the responsibility to push for or conduct research be 

placed on those imposing the risks (and most likely benefitting from the activity) rather than the 
Page 42 Examination of EPA Risk Assessment Principles and Practices 

general public or the population at risk, which may not have the resources to promote such 

research. 

3.5	 Inherent Variability in Biological Response 

3.5.1	 Why Consider Populations and Life-Stages? 

Consideration of the variability among humans is a critical aspect of risk assessment. It is 

the goal of EPA risk assessments to identify all potentially affected populations, including human 

populations (e.g., gender, nutritional status, genetic predisposition) and life-stages (e.g., 

childhood, pregnancy, old age) that may be more susceptible to toxic effects or are highly or 

disproportionately exposed (e.g., children, ethnic groups; see USEPA, 1992a, 2000a, 2002c). 

While susceptible ecosystems and ecological entities (e.g., endangered species) are also 

considered in ecological assessments, the following discussion focuses on human populations 

and life-stages. 

The term “susceptible” is also used to describe sensitive populations/life-stages, as these 

two terms are often used interchangeably and no convention for their use is widely accepted. 

Although “susceptible” has been used to describe susceptibility to toxic effect(s) and 

disproportionate or unique exposures, it is more transparent to identify the two issues separately 

in risk assessment. 

The NRC recognized that many types of variability enter into risk assessment, including 

the nature and intensity of exposure and susceptibility to toxic insult as affected by age, lifestyle, 

genetic background, ethnicity, and other factors (NRC, 1994). The NRC outlined several 

recommendations for considering such variability in risk assessment. The NRC also made 

recommendations specific to risk assessments for children and infants as a potentially susceptible 

group (NRC, 1993). 

The consideration of populations and life-stages has also been described through the 

concept of vulnerability, or the propensity for the system to suffer harm (USEPA, 2003b). 

Factors included in vulnerability are: 

a)	 Susceptibility or sensitivity to adverse effect: increased likelihood of sustaining an 

adverse effect as a relationship to a factor describing the population (e.g., genetic 

polymorphisms, prior immune reactions, disease state, or prior damage) and/or 

life-stage. 

b)	 Differential exposure: differences in current exposure, historical body burden, and 

background exposure. 
Examination of EPA Risk Assessment Principles and Practices Page 43 

c)	 Differential preparedness to withstand a stressor: linked to coping systems and 

resources of an individual, population, or community with respect to prospective 

mitigation efforts (e.g., immunization). 

d)	 Differential ability to recover: linked to coping systems (e.g., differential survival 

rates). 

Legislation has called on EPA to consider potentially susceptible populations and lifestages. The SDWA Amendments mandate that EPA consider risks to groups within the general 

population that are identified as being at greater risk of adverse health effects, including children, 

the elderly, and people with serious illness (SDWAA, 1996). Similarly, the Food Quality 

Protection Act (FQPA) contains special provisions for the consideration of pesticide risks to 

children (FQPA, 1996). In addition to legislative mandates, EPA has further guidance in 

considering health and safety risks to children (EO 13045, 1997; USEPA, 1995b). 

3.5.2	 Are Certain Populations and Life-Stages Always at Greater Risk? 

When conducting risk assessments, EPA examines populations and life-stages that may 

be especially sensitive to the stressor(s) being assessed. This does not imply that these examined 

populations and life-stages are always at greatest risk. For each stressor and exposure scenario, 

different data are available such that there is not a single or exact method for examining potential 

susceptibility and associated risk. Risk assessment often uses an iterative approach. Populations 

and life-stages may be assessed using defaults and assumptions in screening-level assessments, 

while more detailed analyses of these groups will be performed for more refined assessments. 

While the question “Are certain populations and life-stages at greater risk?” needs to be asked in 

all iterations of risk assessment, the answer may not remain the same throughout. 

3.5.3	 How Are Sensitivities to Toxic Effects Considered? 

When data are available to describe the toxicological differences for a susceptible 

population or life-stage, then those data are summarized and analyzed, and the decisions based 

on this information are presented. It is preferable to have population- and chemical-specific data 

to describe a susceptibility to toxic effects. For example, the IRIS RfC for beryllium is based on 

the human subpopulation that is susceptible to chronic beryllium disease (USEPA, 1998b; also 

see section 3.5.5 below). Similarly, if data are available to indicate that susceptible populations 

or life-stages are not at risk, those data are also used. 

When one is analyzing and describing the available toxicity and effects data for 

susceptible populations or life-stages, pragmatic considerations include: 
Page 44 Examination of EPA Risk Assessment Principles and Practices 

a)	 Are epidemiologic, toxicologic, mechanistic, or other studies available that 

investigate reproductive effects, developmental effects or windows of 

susceptibility, multi-generation effects, gender or race differences, genetic 

predisposition, modulation of existing disease, old age, or other possible 

susceptibilities? Were qualitative (i.e., different effect) or quantitative (i.e., same 

effect at lower dose or more severe effect at same dose) differences in 

susceptibility observed? Both? 

b)	 Have the available studies examined the likely and plausible endpoints? 

c)	 Has a distribution been used in assessing toxicity (e.g., benchmark dose lower 

confidence level)? Is it feasible that an entirely different distribution exists for 

susceptible populations or life-stages, such that selecting a central value or even 

tail of the existing distribution is not necessarily representative of susceptibility? 

d)	 How have the findings of available studies been incorporated in the dose-response 

assessment? Is it clear how susceptible populations and life-stages have been 

considered in the quantitative assessment of toxicity? 

e)	 What data are absent/needed? Can information from existing studies be 

synthesized to identify specific data gaps? 

It has been recognized that limited data are currently available for the a priori 

identification of susceptible populations and life-stages for many chemicals and risk assessments 

(USEPA, 2002c). In these situations, it is important that risk assessments clearly identify and 

summarize data needs and uncertainties, in addition to the available data. Typically, when data 

are limited, default practices are used: 

a)	 Non-cancer effects: An intraspecies UF is used to account for variations in 

susceptibility within the human population (USEPA, 2002c). This UF typically 

has a value of 10-fold, but can be increased or reduced when sufficient data are 

available. One can apply the same UF to carcinogens using a non-linear doseresponse model. For example, the IRIS chloroform oral carcinogenesis 

assessment considers the non-cancer assessment to be protective against cancer 

risk, and the same intraspecies UF is applied in the chloroform oral cancer and 

non-cancer assessments (USEPA, 2001c). 

A database UF may also be applied for deficiencies in the available data or when 

existing data suggest that additional data may yield a lower reference value 

(USEPA, 2002c). This UF is most often used when developmental or twogeneration reproduction studies are not available, but it may be applied in other 
Examination of EPA Risk Assessment Principles and Practices Page 45 

situations to account for the lack of data for potentially susceptible populations or 

life-stages. 

b)	 Cancer effects: An evaluation should be made as to whether low-dose linear 

extrapolation is sufficient to protect susceptible populations (USEPA, 2003c). 

For example, available data indicate that early life exposure to mutagenic 

carcinogens may lead to greater incidence of cancer than the same lifetime 

average daily dose received later in life. This information has been reviewed by 

EPA and published in draft (USEPA, 2003d). 

3.5.4	 Are Unique or Disproportionate Exposures Considered? 

Unique or disproportionate exposures are considered in parallel to the toxicity 

considerations described in the preceding section. Available data describing potential unique or 

disproportionate exposure are summarized and decisions based on this information are presented. 

Again, it is preferable to have population- and chemical-specific data to address specific 

exposure scenarios. For example, the Integrated Exposure Uptake Biokinetic Model (IEUBK) is 

a very specialized tool that allows detailed analysis of lead exposures for children ages 6 months 

to 7 years (USEPA, 2001d). 

While some risk assessments consider exposure to the population as a whole, many only 

consider exposed populations, not unexposed members of the general population. For example, 

a hazardous waste assessment only considers the people who live near or on the site, or near the 

incinerator or point source involved. 

For exposure assessment, pragmatic considerations for unique and disproportionate 

exposures include: 

a)	 Are there unique or disproportionate exposures based on age, race, ethnicity, 

gender, lifestyle, cultural practices, economic status, or other considerations? 

b)	 How well do existing studies or distributions of exposure consider identified 

populations or life-stages? 

c)	 Have the exposures within these groups been considered in detail? For example: 

pica soil ingestion by children, consumption of breast milk by infants, traditional 

or subsistence diets, mouthing of objects by toddlers, different/increased food and 

water consumption rates for children, restrictions on mobility (e.g., children 

cannot leave a location or residence by choice as adults can; elderly populations 

often have reduced mobility). 
Page 46 Examination of EPA Risk Assessment Principles and Practices 

d)	 Have unique or disproportionate childhood exposures been integrated for lifetime 

exposure? Children are not children for a chronic duration, but all adults spent a 

portion of their lifetime in childhood; this needs to be taken into account as a 

source of unique and disproportionate exposure as a part of a continuous lifetime. 

When exposure data for certain populations and life-stages are limited, exposure 

assumptions and default values are used to assess plausible current and/or future exposure 

scenarios. The exposure factors published in EPA’s Exposure Factors Handbook are widely 

used in Agency exposure assessments (USEPA, 1997d, 2002d). In some scenarios, additional 

default data may be used to describe situations the handbook does not cover. For example, 

exposure parameters used to estimate childhood exposures due to mouthing exposures have been 

used in pesticide exposure assessments (USEPA, 1997d). 

It is essential that risk assessments, and exposure assessment in particular, explicitly 

address environmental justice concerns. The goal of environmental justice is to ensure that all 

people, regardless of race, national origin, or income, are protected from disproportionate 

impacts of environmental hazards (EO 12898, 1994). For example, concerns regarding risks 

associated with subsistence and the practice of traditional lifeways have been voiced by Native 

American people across the United States and by the National Environmental Justice Advisory 

Council (NEJAC, 2002). Executive Order 13175, “Consultation and Coordination With Indian 

Tribal Governments,” requires EPA to consult with tribes regarding protection of their land and 

health. Tribes have specifically requested that EPA enter into formal consultation with them on 

environmental issues, including exposures unique to tribes. These exposures are a result of 

cultural or behavioral preferences of certain native populations. EPA typically bases its decisions 

on the general U.S. population; statistics derived from the general population’s exposure patterns 

do not necessarily represent exposure patterns for all peoples. For example, the exposure 

distribution for four Columbia River Treaty Tribes (CRITFC, 1994; also, see table 3-1, below) 

shows that these peoples’ average fish consumption rate is much higher than the rate for the 

average U.S. person. This disproportionate exposure leaves Native Americans in these 

communities with exposures that are underestimated or ignored in risk assessments. 
Examination of EPA Risk Assessment Principles and Practices Page 47 

Table 3-1: Fish Consumption Patterns for General U.S. and Native American Populations 

Exposure Parameter 

General Public Columbia River Treaty Tribes 

AFC HFC AFC HFC 

Tissue chemical concentration Average Average Average Average 

Ingestion rate of fish tissue (g/day) 

Adults 7.5a 142.4b 63.2c 389d 

Children under 15 2.83a 77.95b — — 

Children under 6 — — 24.8c 162d 

AFC — average fish consumption; HFC — high fish consumption 

a

 Mean U.S. per capita consumption rate of uncooked freshwater and estuarine fish (USEPA, 2000f). 

b

 99th percentile U.S. per capita consumption rate of uncooked freshwater and estuarine fish (USEPA, 2000f). c

 Mean consumption rate for fish consumers in the Umatilla, Nez Perce, Yakama, and Warm Springs Tribes of the Columbia 

River Basin (CRITFC, 1994). 

d

 99th percentile consumption rate for fish consumers in the Umatilla, Nez Perce, Yakama, and Warm Springs Tribes of the 

Columbia River Basin (CRITFC, 1994). 

Situations for unique exposures also need to be considered. For example, the Asian 

Pacific American population typically consumes foods that are not included in the U.S. diet 

survey (Sechena et al., 2003). Thus, their exposure to contaminants in seaweed or moon snails 

are not included in routine market basket surveys or the resulting exposure estimates. 

Sometimes disproportionate exposures occur because of racial or socioeconomic 

conditions. For example, exposure to environmental and household agents can trigger asthma 

attacks (AAP, 2003); the Centers for Disease Control and Prevention reported that asthma 

mortality rates among African Americans are 2.5-fold higher than among Caucasians (CDC, 

2003). Similarly, children of low income families constitute 83% of the children ages 1 to 5 with 

elevated blood lead levels that may result in health effects. 

3.5.5 Is Human Variability Considered in Occupational Cohorts? 

Occupational epidemiologic cohorts are an important source of human toxicity 

information for use in risk assessment. Occupational studies are conducted with select 

populations. Because of their presence in the workforce, these persons are believed to be healthy 

(this is the “healthy worker effect”), predominately male, and generally adult. These factors 

obviously limit a population’s diversity with respect to susceptibility. As a result, the general 

approach is to use an intraspecies UF greater than 1 for non-cancer endpoints when the RfD or 

RfC is based on occupational data. This practice is consistent with the recommendation that the 

reduction of the intraspecies UF of 10-fold be considered only if available data are sufficiently 

representative of the exposure/dose-response data for the most sensitive populations and lifestages (USEPA, 2002c). 
Page 48 Examination of EPA Risk Assessment Principles and Practices 

The 2003 IRIS database was examined for all instances in which human occupational 

data had been used to set an RfD and/or RfC. In the six instances in which occupational data had 

been used to set the RfD, an intraspecies factor of 10 was used three times, an intraspecies factor 

of 3 was used twice, and an intraspecies factor of 1 was used once. In the latter case (that of 

beryllium and beryllium compounds), the health effect involved is beryllium sensitization and 

progress to CBD. The intraspecies factor of 1 was chosen because only a small percentage of the 

population is sensitive to CBD and sensitization: the affected population already represented a 

sensitive group of people. 

Occupational epidemiologic cohorts are also used in cancer assessments. For example, in 

the health assessment of 1,3-butadiene, the cancer slope factor (CSF) was derived using an 

occupationally exposed male cohort. In animal studies, mammary gland cancer was the only 

common tumor in mice and rats (USEPA, 2002e), so a factor of 2 was used to protect the female 

population. This is an example of observations from the occupational cohort — which did not 

include women and children — not necessarily representing the risks to the larger, more diverse 

human population. 

3.5.6	 What Needs To Be Done To Address the Risk Assessment of Populations and 

Life-Stages? 

Risk characterization is the opportunity to bring together all population and life-stage 

considerations, including the toxicity and exposure concerns discussed in the preceding sections. 

A risk characterization considers all information, integrated through a logical and transparent 

approach, and provides a thorough discussion of the findings and related uncertainties (USEPA, 

2000a). A risk characterization needs to provide a clear answer to the question, “Are there 

susceptible populations and life-stages?” 

Overarching guidance is needed to improve the ease and consistency of risk assessment 

for susceptible populations and life-stages. Such guidance needs to provide more detail, 

methods, and advice for considering the issues discussed in this section. As a part of such 

efforts, EPA needs to continue developing a strategy and methodologies to address the 

cumulative risks of unique or disproportionate exposures. 

Continued research and the availability of new types of information (e.g., genomics, 

improved early life animal toxicity testing, cultural practices awareness, demographics) will 

provide an improved ability to identify and describe susceptible populations and life-stages. It is 

anticipated that these advances will improve the quality of risk assessments and guidance 

documents — both for risk assessment in general and for susceptible groups in particular. Until 

such data are available, the current practice of considering available data and following Agency 

practice where data are limited is a reasonable approach. The Agency continues to strive to be 
Examination of EPA Risk Assessment Principles and Practices Page 49 

clear and transparent in explaining the assessment of populations, uncertainties, and life-stages in 

risk assessment. 

3.6	 What Is EPA Doing To Improve Assessment of Uncertainty and Variability? 

3.6.1	 How Can EPA Move Forward Toward Quantitative Characterization of 

Uncertainty in Dose-Response? 

In successive versions of its Cancer Guidelines, EPA expressed increasing emphasis on a 

full examination of uncertainties, with the recognition that both qualitative and quantitative 

approaches to uncertainty assessment are important and can (applied appropriately) help clarify 

the nature of assessment findings. The use of sophisticated uncertainty tools also involves 

substantial issues of science and mathematics, as well as specialized issues such as the 

appropriate presentation and characterization of probabilistic estimates in the decision making 

context where appropriate. We note that active research is underway in this field, and that EPA 

guidelines are intended to be flexible to allow use of advances in this field as they develop. As 

also discussed in various places in the guidelines, substantial quantitative uncertainty analyses 

are also appropriate (and at a more advanced stage of development) for a number of important 

components of cancer risk assessments, including exposure assessment, characterization of 

statistical uncertainty in fits of both descriptive and biologically based dose-response models, and 

dosimetry estimates using pharmacokinetic data and modeling. It is not, however, EPA’s intent 

to suggest that full probabilistic models of cancer risks are generally feasible at this time, or that 

the role of a qualitative presentation of uncertainties should be diminished. 

EPA health scientists currently conduct research that both generates necessary 

quantitative data and integrates data-driven hypotheses on mode of action into biologically based 

models. These models simulate key biological processes and provide quantitative predictions 

that improve risk assessment, including cross-route extrapolation and animal-to-human 

extrapolation. One can use these models to develop better estimates of variability in the human 

population, to identify susceptible subpopulations, to help determine the applicability of 

proposed modes of action, and to derive likelihoods of adverse effects in humans. 

3.6.2	 Will EPA Move Toward Integrated Uncertainty Analysis? 

EPA recognizes the importance of uncertainty analysis and the value of having an 

integrated framework for evaluating uncertainty within an overall analysis, rather than as an a 

posteriori exercise that generates deterministic estimates. EPA is working toward integrated 

probabilistic frameworks within risk models now being developed. Once these are implemented, 

EPA will be able to conduct probabilistic analyses as part of any original analysis, though it is 

recognized that probabilistic frameworks will be pointless without adequate data to insert. 
Page 50 Examination of EPA Risk Assessment Principles and Practices 

Models that include integrated probabilistic frameworks include APEX, TRIM, and 

CALENDEX. 

In addition, EPA is developing a flexible modeling framework, the Multimedia Integrated 

Modeling System (MIMS), which provides an infrastructure for modeling that supports 

composing, configuring, applying, and evaluating models. It can support connecting models in a 

controlled manner, provide graphical user interfaces to help users configure models, and manage 

the execution of complex systems of models. MIMS has been designed to support a wide range 

of models. One of the benefits of using MIMS to manage model executions is that it can initiate 

repetitive executions, including simulation of multiple scenarios and sensitivity, uncertainty, and 

optimization studies. 

MIMS’s sensitivity and uncertainty capabilities are designed to work with a broad range 

of models or modeling systems. The initial sensitivity capability will support local evaluation. 

Uncertainty capabilities will initially include simple Monte Carlo and Latin Hypercube 

Sampling. More sophisticated capabilities, such as analysis of variance for sensitivity studies, 

are planned for the future. In addition to handling parameter sampling and model execution, 

MIMS will provide data analysis tools to help the user interpret the results of the computations. 
Examination of EPA Risk Assessment Principles and Practices Page 51 

4. CONSIDERING INFORMATION GAPS IN HEALTH 

ASSESSMENTS: USE OF DEFAULT AND EXTRAPOLATION 

ASSUMPTIONS 

4.1 Default Assumptions 

Given the nature of uncertainty and data gaps, it is accepted practice to use default 

assumptions (sometimes simply called defaults) to address these uncertainties. A default 

assumption is “the option chosen on the basis of risk assessment policy that appears to be the 

best choice in the absence of data to the contrary” (NRC, 1983). The NRC, in its review of EPA 

risk assessment practices titled Science and Judgment in Risk Assessment (NRC, 1994), 

supported EPA’s use of defaults as a reasonable way to deal with uncertainty. That report stated 

that EPA should have principles for choosing default options and for judging when and how to 

depart from them. It identified a number of criteria that NRC believed ought to be taken into 

account in formulating such principles, including protecting the public health, ensuring scientific 

validity, minimizing serious errors in estimating risks, maximizing incentives for research, 

creating an orderly and predictable process, and fostering openness and trustworthiness. 

The first two sections (4.1.1 and 4.1.2) discuss default assumptions in general. The rest 

of section 4.1 discusses specific examples of defaults that have been highlighted in the comments 

we received. 

4.1.1 How Does EPA Use Default Assumptions? 

EPA’s current practice is to examine all relevant and available data first when performing 

a risk assessment. When the chemical- and/or site-specific data are unavailable (i.e., when there 

are data gaps) or insufficient to estimate parameters or resolve paradigms, EPA uses a default 

assumption in order to continue with the risk assessment. Under this practice EPA invokes 

defaults only after the data are determined to be not usable at that point in the assessment — this 

is a different approach from choosing defaults first and then using data to depart from them. The 

default assumptions are not chemical- or site-specific, but are relevant to the data gap in the risk 

assessment. They are based on peer reviewed studies and extrapolation to address specific data 

gaps. These defaults are based on published studies, empirical observations, extrapolation from 

related observations, and/or scientific theory. 

EPA’s use of defaults is appropriate. These choices are well within the range of plausible 

outcomes and often at specific percentiles (for variability) within that range of observation. The 

aim of the risk assessment is to ensure that EPA does not underestimate risk to a chemical or 

stressor; the default assumptions used in the risk assessment are used to help pursue this goal. 
Page 52 Examination of EPA Risk Assessment Principles and Practices 

Conversely, the intent of using default assumptions is not to produce a risk assessment that 

appreciably overestimates the risk being assessed. 

A default is usually based on some data and/or some consensus agreement for the generic 

situation pertinent to the data gap being assessed. Defaults usually have undergone peer review, 

in which interested parties are involved or have the opportunity to be involved in those reviews. 

For example, many of the exposure defaults we use are found in EPA’s Exposure Factors 

Handbook, a resource that was peer reviewed and continues to be examined and re-evaluated. 

When chemical- or site-specific information becomes available and is adequate to use, our risk 

assessments attempt to use those data rather than the default(s) — assuming those data are 

relevant for the specific risk being assessed, and the peer review supports their use. 

There is always room for improvement in highlighting where these defaults are used in 

actual assessments, and one should not assume that people who read an assessment will 

automatically know that a default position is used (NRC, 1994). Where and when we use default 

assumptions to address a data gap may not be clear to many. We need to point to the places 

where the default was developed and explain why it is a reasonable assumption. 

4.1.2 Are Default Assumptions Science Policy?

 One of the major comments EPA received suggests that default assumptions are actually 

conservative policy decisions embedded in risk assessments. While default assumptions have 

data and/or scientific consensus supporting them, use of defaults to address data gaps in a risk 

assessment is a science policy decision. In keeping with the EPA’s goal of protecting public 

health and the environment, the default assumptions are used ensure that we do not 

underestimate risk to chemicals and stressors. (Again, though, they are not intended to overtly 

overestimate a risk.) Defaults are ideally developed in a transparent manner, and we believe this 

is the practice of the Agency. Therefore, while defaults are indeed used in risk assessments, they 

are not buried (hidden) policy decisions. Recently, the GAO reflected that we use assumptions 

(defaults) as an unavoidable part of risk assessment because science cannot always provide clear 

answers to questions at various stages of an assessment (USGAO, 2001). Particular assumptions 

are chosen through an evaluation of available scientific information or policy decisions related to 

our regulatory mission and mandates. Due to the complexity of risk assessments, transparency in 

choice and use of assumptions is critical in risk assessments and communication efforts. 

If a default assumption is a policy position that some deem too conservative or not 

conservative enough, this issue can be addressed during the peer review. Generally speaking, 

EPA’s current default assumptions are positions based on data or scientific consensus and 

supported through peer review. Some older risk assessments may contain defaults that have not 

yet been reexamined in light of new data. Thus, in the context of current data, they appear 

inadequate (i.e., to some, they can be too conservative or not conservative enough). This is not 
Examination of EPA Risk Assessment Principles and Practices Page 53 

an inherent problem with the original risk assessments, but a process problem for EPA: EPA 

needs to determine the need to reexamine older assessments based on current priorities, needs, 

and resources. 

4.1.3	 Does Any Change Seen in Animals Indicate There Will Be a Problem for 

Humans? 

It is generally accepted that there can be numerous changes to the recipient organism (the 

animal under study) following exposure to a chemical, some of which may be beneficial, 

adaptive, early manifestations on a continuum to toxicity, overtly toxic, or several of these things 

in combination. Unless there are data to indicate otherwise, a change that is considered adverse 

(i.e., associated with toxicity) is assumed to indicate a problem for humans. 

It is recognized that a diversity of opinion exists regarding what is “adverse” versus 

“adaptive,” both within EPA and in the general scientific community. At present, there is no 

Agency-wide guidance from which all health assessors can draw when making a judgment about 

adversity. Therefore, various experts may have differing opinions on what constitutes an adverse 

effect for some changes. Moreover, as the purpose of a risk assessment is to identify risk (harm, 

adverse effect, etc.), effects that appear to be adaptive, non-adverse, or beneficial may not be 

mentioned. 

As a further look at this issue, an “adaptive” example is used. The human body is capable 

of adapting to certain toxic insults. When adaptive responses become adverse and irreversible is 

not yet defined. In cases where data are not available to determine when the capacities of repair 

mechanisms are exceeded and adaptive responses become toxic, health assessments are based on 

any adverse response that is deemed biologically significant. As a general principle, our practice 

is not to base risk assessments on adaptive, non-adverse, or beneficial events. 

The IRIS database includes some examples of effects reported at the designated NOAEL 

level, but their toxicological significance was questioned and not relied upon in setting 

RfDs/RfCs; for example, mild nasal lesions and hyaline droplet degeneration in the nasal cavity. 

These effects were not considered of sufficient toxicological significance to warrant designating 

the levels that produced them as LOAELs. In these instances, the LOAELs were set at the next 

highest dose level, based on more severe nasal responses. 

There are also differences of opinion regarding the severity of effects and their qualitative 

and quantitative relationship to more overt toxicity. To explore past Agency practice, we 

performed a series of IRIS database searches on a selected example: “liver weight” as the sole 

critical effect, expanded slightly to incorporate examples in which liver weight appears to be 

listed as the predominant critical effect, along with other effects. The results indicate that liver 

weight change has been listed as a sole critical effect in the IRIS database in seven profiles (and 
Page 54 Examination of EPA Risk Assessment Principles and Practices 

as a predominant effect in three more; examples dating from 1987 to 1994). More in-depth 

examination of these IRIS records generally indicates that at least one additional toxicity 

consideration was incorporated into the Agency decision at that time. This additional toxicity 

information comes in three forms: (1) listing of co-critical adverse events; (2) supporting 

information on adverse effects from other studies, at lower, similar, or slightly higher doses, that 

have not been reported in the IRIS critical events column; and/or (3) a clear progression of 

toxicity at slightly higher doses. All three sources of additional toxicity information may be 

discussed in a particular IRIS file, with the predominant information being supportive toxicity 

data at similar doses in other studies. The extent to which additional toxicity information is 

available and/or reported varies among the chemical assessments. This examination does not 

rule out the possibility of using liver weight (or other changes) as a critical effect if it is deemed 

to contribute to adversity in an assessment. 

4.1.4	 Are Benign (Histologically Non-Malignant) Tumors Presumed To Have 

Potential To Progress to Malignant Tumors, and Are They Counted as if 

They Were Malignant? 

When evaluating the results of long-term bioassays in rodents for evidence of potential 

carcinogenicity, we do not always presume that all benign tumor types have the potential to 

progress to malignant tumors. Although the total incidence of benign tumors is considered in 

evaluations of carcinogenic potential, the weight placed on evidence of an increase in a benign 

tumor response is subject to consideration of the potential to progress to a malignant tumor. 

For example, a uterine polyp, a benign tumor, is not considered to have the potential to 

progress to a malignant tumor; in the absence of evidence of an increase in malignant tumors in 

the same tissue (or in other tissues), evidence of carcinogenic potential would be considered to be 

weak. In such cases, the chemical would likely be assigned the descriptor “Suggestive Evidence 

of Carcinogenicity, but Not Sufficient To Assess Human Carcinogenic Potential” or “Not Likely 

To Be Carcinogenic to Humans” and a quantitative risk assessment would not be recommended 

(USEPA, 1999b). 

On the other hand, hepatocellular adenomas are presumed to have the potential to 

progress to hepatocellular carcinomas. This is because there is considerable evidence showing 

that hepatocellular adenomas may progress to carcinomas in both rodents and humans. In 

addition, there is no clear histological demarcation between hepatocellular adenomas and 

carcinomas. Thus, when the results of rodent bioassays give evidence of the induction of both 

adenomas and carcinomas, the combined incidence of the two tumor types is considered in a 

weight-of-evidence evaluation of potential carcinogenicity. This approach is consistent with that 

of the National Toxicology Program (NTP), which recommends combining hepatocellular 

adenomas and hepatocellular carcinomas for an overall analysis of carcinogenicity (in addition to 

analyzing them separately). 
Examination of EPA Risk Assessment Principles and Practices Page 55 

4.1.5	 Should We Expect Target Organ Concordance Between Effects Observed in 

Animals and Those Expected in Humans? 

Thus far, there is evidence that growth control mechanisms at the level of the cell are 

homologous among mammals, but there is no evidence that these mechanisms are necessarily 

target organ concordant. For example, Haseman and Lockhart (1993) examined a database of 

379 long-term carcinogenicity studies in rats and mice to evaluate sex and species correlations in 

site-specific carcinogenic responses. Within a species, most target sites showed a strong 

correlation between males and females. If a chemical produced a site-specific carcinogenic effect 

in female rats or mice, there was a 65% probability that it would be carcinogenic at the same site 

in males. On the other hand, the interspecies correlation was lower. Approximately 36% of the 

site-specific carcinogenic effects observed in one species were also observed in the other species. 

EPA’s cancer risk assessment guidelines recommend that site concordance of tumor 

effects between animals and humans be considered in each case. Site concordance is not 

assumed a priori. On the other hand, in establishing a mode of action that involves certain 

processes with consequences for particular tissue sites (e.g., disruption of thyroid function, 

hormonal effects leading to mammary tumors), one would expect site concordance when 

evaluating the weight of evidence. We address the issue of target organ concordance on a caseby-case basis, considering information on the mode of action. For example, based on a 

significant body of data, EPA determined that male rat kidney tumors are not relevant to the 

assessment of human cancer risk where tumors arise as a result of accumulation of a protein 

(alpha-2u-globulin) unique to the male rat (USEPA, 1991a). Similarly, the reproductive and 

developmental endpoints need to be judged considering what is known about the mode of action 

for an individual chemical as well as how the effects in animals may or may not be predictive of 

effects in humans. 

Similarly, the expectation of concordance of fetal effects would depend on knowledge 

about mode of action. Most fetal effects are exquisitely dependent on time as well as dose. As 

the developmental process of laboratory animals and humans differs, especially for rodents, 

slight differences in the pattern of exposure or the toxicokinetics in the mother may greatly affect 

the observed outcome. In this regard, it is worth noting that the standard metric of daily dosing 

may have different implications for a laboratory animal than a human. Moreover, certain 

reproductive endpoints that are useful for evaluating fetal effects in rodents have no human 

equivalent, e.g., number and percent of live offspring per litter. With regard to fetal effects, for 

example, we consider both individual effects and total affected implants. Changes in dose may 

not only change the specific malformation, but also cause fetolethality rather than malformation. 

Thus, concordance of endpoint may not be the best predictor of developmental effects in people. 
Page 56 Examination of EPA Risk Assessment Principles and Practices 

4.2	 Extrapolations 

Some federal regulatory agencies can estimate risks based on historical data about those 

risks. Thus, we have information on risks such as house fires, traffic accidents, and occupational 

injuries. At EPA, we are often faced with the problem of estimating the risk from chemicals for 

which we have limited or no historical data, e.g., new chemicals. Even when such a chemical 

has been in use, most if not all of the available information about its toxicity comes from studies 

with laboratory animals that use concentrations that are designed to produce effects. Similarly, 

available human data usually involve either exposures at high concentrations for short durations 

(e.g., accidents) or occupational exposures. Our concerns often involve situations where people 

might be exposed to much lower concentrations of the chemical for various periods of time. We, 

therefore, need to use procedures to extrapolate from responses (or lack thereof) at exposures 

where data are available to exposures of interest. Frequently, we also need to extrapolate from 

responses of laboratory animals to those that might reasonably be expected in people. 

The rest of section 4.2 discusses how EPA extrapolates from animals to humans. 

4.2.1	 Why Does EPA Use Experiments With Animals To Predict Effects in 

Humans? 

As a general rule, we assume that toxic responses observed in laboratory animals are 

indicative of toxic responses that are likely to occur in people. To predict responses in people 

more accurately, we would prefer information from animals that are as similar to people as 

possible. The use of other mammals, such as dogs, rats, and mice, as models for responses in 

humans is based on the assumption that there are important similarities among mammals in the 

way they respond to chemicals. A qualitative similarity has been established in the response of 

laboratory animals and humans to carcinogenic substances. For example, one analysis of 

chemicals that have been tested by the NTP for their ability to produce cancer (Huff et al., 1991) 

indicates that rats and mice predict responses for each better than 80% of the time for liver 

cancers. Most known human carcinogens have been shown to be positive for tumorigenicity in 

well-conducted animal studies. Data are not available to determine whether the opposite is true. 

Nevertheless, important differences can exist. Metabolic patterns of both activation and 

deactivation along with differences in pharmacokinetics, for example, can give rise to significant 

differences in sensitivity between species and within species. Unless we know the rates of the 

activation-deactivation processes, for example, it is impossible to predict the differences in 

response among species. Presently, there is a paucity of comparative data on metabolism for 

specific chemicals and other interspecies differences that can affect toxicity. Important 

differences require experiments on the toxicokinetics or toxicodynamics within each species. 

We cannot be certain which results are most like those for humans in the absence of these data. 
Examination of EPA Risk Assessment Principles and Practices Page 57 

The increasing use of mode of action information will make it clearer whether a similarity or 

difference is apparent between test animals and humans. 

4.2.2	 How Does EPA Adjust for the Differences Between Animals and Humans? 

If we have sufficient information about how the environmental agent is absorbed, 

distributed, metabolized, and excreted in the laboratory animal and the human and about the 

physiologies of the two species, we can develop a model (such as a PBPK model) that will allow 

us to compare how differences in these factors and processes affect the biologically effective 

dose to a target tissue. For some chemicals, this process has resulted in a very specific 

adjustment from the animal to the human. When information is limited, we can only make 

general adjustments. 

Our practice is to interpret the findings of long-term rodent bioassays in conjunction with 

results of subchronic studies along with toxicokinetic studies and other pertinent information, if 

available. When data are appropriate, we use metabolic and toxicokinetic data to: 

a)	 Identify and compare the relative activities of metabolic pathways in animals and 

humans and at different ages. 

b)	 Describe anticipated distribution within the body and possibly identify target 

organs. 

c)	 Identify changes in toxicokinetics and metabolic pathways with changes in dose. 

d)	 Determine the bioavailability via different routes of exposure by analyzing uptake 

processes under various exposure conditions. 

This issue is addressed on a case-by-case basis where data are available. When speciescomparative data are not available, we assume concordance between the animals tested and 

humans. 

4.2.3	 Does EPA Use the Most Sensitive Animal Results To Predict Effects in 

Humans? 

Use of the most sensitive species, strain, sex, age, etc., was originally justified on the 

need to be health protective in the absence of data. From the 1970s through the 1990s, the most 

sensitive responding species (given several data sets to choose from) were frequently selected: 

very little information was available as to what was scientifically the most representative choice 

for human risk prediction. Combined with UFs and other upper-bound estimates, basing cancer 

and non-cancer risks on the most sensitive animal data gives reasonable assurance that the 
Page 58 Examination of EPA Risk Assessment Principles and Practices 

potential for harm will not be underestimated, most likely even when some toxicity endpoints 

have not been evaluated. Our early cancer risk assessment guidelines (USEPA, 1986a) comment 

on the use of the most sensitive animal response in the absence of appropriate human studies. 

The most recent draft guidelines (USEPA, 2003c) suggest a wide range of considerations based 

on knowledge and uncertainty for selecting the appropriate data set(s), assuming several sets are 

available. 

Other federal agencies use a similar approach. According to Rhomberg (1997), the 

Occupational Safety and Health Administration (OSHA) also generally uses the data set 

demonstrating the highest estimated sensitivity, although it tends to “present several alternatives 

together or do several analyses and present the median result.” The Consumer Product Safety 

Commission (CPSC) also uses similar methods, but differs in its approach to combining tumors 

from different sites. 

The availability of more biological knowledge or insight potentially enables the risk 

assessor to make more scientifically informed choices among the available experiments with 

animals for use in dose-response analysis. The recent draft cancer guidelines (USEPA, 2003c) 

ask for a full display of all data options as well as the rationale for why one data set is selected 

and why others are rejected. Data from humans are preferred, although they are not always 

amenable to quantitative risk assessment. Alternatively, data from species that respond most like 

humans should be used if information to support this selection exists. If this is not known, all of 

the available data sets are considered and compared. An informed, scientific judgment is made 

as to what data best represent the observed data and important biological features such as mode 

of action. When faced with critical uncertainties or data gaps, however, we will favor a selection 

that ensures against underestimating risk as a policy choice. 

We are continuing our trend toward using relevant data before using a default assumption 

to address data gaps. Use of toxicokinetic and toxicodynamic data for interspecies extrapolation 

should better address this source of uncertainty. Using the available data should accommodate 

some of the differences among species. For example: 

a)	 When data on mode of action indicate that most the sensitive species/strain/sex is 

not an appropriate model for humans, e.g., alpha-2u-globulin, those data are not 

used for estimating human risks. 

b)	 In some cases, use of the data that estimate the highest potential for harm can 

result in the use of less-well-designed studies, a concern that should be attenuated 

by increased use of the benchmark dose procedure. 

c)	 Additionally, we have begun to investigate the utility of meta-analysis as one way 

to evaluate an entire database quantitatively. 
Examination of EPA Risk Assessment Principles and Practices Page 59 

4.2.4	 Is an Additional Body Scaling Factor for Children Considered? 

The draft final cancer guidelines (USEPA, 2003c) suggest that different adjustments may 

need to be made for children inhaling carcinogenic substances than for adults, but the guidelines 

do not specify how this should be done. For example, their language has been incorrectly 

interpreted as suggesting the use of a scaling factor of body weight to the 0.75 power before 

adjusting for a child’s breathing rate. These draft guidelines recognize that correlations exist 

“between breathing rate, respiratory tract dimensions, and body weight.” 

The methods that could be used for interspecies extrapolation with appropriate 

adjustments for children are the subject of draft guidance (USEPA, 2003d), ongoing discussions, 

and further assessment by the Agency. The supplemental guidance for early-life exposures, for 

example, illustrates the importance of information regarding the mode of action of the carcinogen 

on evaluating the potential risks to children. We understand the need to address this issue further 

and will do so, in part, based on the comments received on the draft guidance. 

4.2.5	 How Are Effects Observed at High Exposures Used To Predict Responses to 

Much Lower Exposures? 

The role of animal tests is to provide maximum detectability of effects within the narrow 

constraints of test sensitivity. This is typically achieved through the use of “high” exposures or 

doses to increase the probability that a possible adverse effect will be observed. This practice is 

usually used when sample sizes are too small to be sensitive to effects that are not overtly toxic 

or seen in response to potent stressors. Generally, the methods used to extrapolate from high to 

lower exposures differ for cancer and non-cancer effects. Both are discussed below. It should be 

noted that as more mode of action information becomes available, the distinction between cancer 

and non-cancer endpoints will dissipate and the dose-response evaluation is more likely to 

distinguish between linear and nonlinear extrapolations from high exposures to lower exposures. 

Cancer 

Carcinogenic effects are typically evaluated using a long-term rodent carcinogenicity 

bioassay. Current practice for these tests suggests using at least 50 animals per sex per dose 

group in each of three treatment groups and in a concurrent control group, usually for 18 to 24 

months, depending on the rodent species tested. One group is exposed to a high dose, often the 

maximum tolerated dose, or MTD (discussed further below). Two groups are given lower doses 

of the test compound. The high dose in long-term studies is generally selected to provide the 

maximum ability to detect treatment-related carcinogenic effects while not compromising the 

outcome of the study through excessive toxicity or inducing inappropriate toxicokinetics (e.g., 

overwhelming absorption or detoxification mechanisms). The purpose of the two or more lower 
Page 60 Examination of EPA Risk Assessment Principles and Practices 

doses is to provide some information on the shape of the dose-response curve, including possible 

subtle, precursor, and/or early events that can provide a more meaningful dose-response 

relationship relative to anticipated human exposures. Similar protocols have been and continue 

to be used by many laboratories worldwide. 

Because of biological variability and the relative low sensitivity of most bioassay designs, 

positive evidence at the high dose is considered a rebuttable presumption of carcinogenicity. 

These findings are not altered by negative data at lower exposures unless there are supporting 

studies that provide a biological basis to conclude that the mode of action responsible for the 

high-dose effects is not relevant at lower doses and therefore for lower levels of exposure 

experienced by humans. 

There is often a need to estimate risks for exposures significantly below exposures for 

which we have data. Two methods are used for this process. For mutagenic chemicals and those 

for which insufficient data exist to determine a mode of action, a point of departure (POD) at the 

low end of the exposures for which we have data is selected and a straight line is drawn from (the 

lower confidence limit on) that dose to zero. A nonlinear approach is selected when there are 

sufficient data to ascertain the mode of action and conclude that it is not linear at low doses and 

the agent does not demonstrate mutagenic or other activity consistent with linearity at low doses. 

The use of low-dose linear assumptions for mutagenic carcinogens has been the peer 

reviewed, common practice for decades for many federal regulatory agencies including EPA. 

Modifications for non-mutagenic carcinogens are currently under review. 

Non-Cancer 

For non-cancer effects, we routinely use the absence of an adverse effect (a NOAEL) in a 

relevant animal study as the basis for estimating a negligible risk exposure level for humans. 

Adjustments are made to the NOAEL to account for limitations in knowledge, i.e., uncertainty, 

or limitations inherent in the test system, e.g., the use of relatively small numbers of genetically 

relatively homogeneous animals. Similarly, mode of action, physiological differences, or other 

factors have been used to improve the accuracy of estimation of the RfD or RfC for non-cancer 

endpoints. 

The Agency has also been using and is continuing to refine the BMD methodology 

(USEPA, 2000g) for estimating non-cancer risks. In this procedure, the POD is derived from the 

dose-response curve, and further adjustments analogous to those for the NOAEL may be 

necessary to account for similar limitations of the data. 

The EPA RfD/RfC review (USEPA, 2002c) delineates several options that have been 

used as the Agency developed more accurate methods for evaluating these parameters from 
Examination of EPA Risk Assessment Principles and Practices Page 61 

responses in animals. It also provides a series of examples, as well as lists of recommendations 

that address the predictivity of results from animal studies. 

4.2.6	 Why Is the Maximum Tolerated Dose Used? 

Because tests for cancer are long and expensive, they are designed to increase the 

likelihood of detecting a positive response with a limited number of animals. The highest dose 

that does not produce other overt toxicity (i.e., the MTD) and fractions of that dose are selected 

for the dosing regimen. The MTD is usually determined from a shorter, e.g., 90-day, experiment 

with the same animal species and chemical. 

The use of such high doses in animal cancer tests has been the subject of considerable 

debate. Limitations inherent in using the MTD approach and suggestions for improvement have 

been controversial since its use became standard. Testing at the MTD has been criticized 

because it may cause damage which may lead to cellular proliferation, increased mitosis, and 

eventually carcinogenicity. It has been argued that biochemical and physiological distortions 

occurring at high doses may lead to toxicity-induced carcinogenic effects that might not be 

expected at lower doses (Ames and Gold, 1990). Haseman (1985) has shown that more than 

two-thirds of the carcinogenic effects detected in feeding studies conducted under the NTP would 

have been missed if the highest dose had been restricted to doses below the MTD, if the 

chemicals were, in fact, animal carcinogens. Additionally, Bickis and Krewski (1989) found the 

upper confidence limit on the linear term (q1

*

) from the linearized, multi-stage model based on 

263 data sets to be highly correlated with the maximum dose tested. 

The draft final cancer guidelines (USEPA, 2003c) cautions the assessor in the use of 

results from the MTD and recommends that this issue be addressed on a case-by-case basis in the 

context of other study results and other lines of evidence. The guidelines state that the results of 

such studies would not be considered suitable for dose-response extrapolation if it is determined 

that the mode of action underlying the tumorigenic response at high doses is not operative at 

lower doses. For example, EPA developed a science policy position for agents that cause thyroid 

follicular cell tumors as a result of high-dose effects that are not present at low doses (USEPA, 

1998c). In addition, the recent chloroform risk assessment was predicated on the recognition of a 

high-dose effect leading to cell proliferation and then carcinogenicity. The MCL was therefore 

developed based on an assumption of nonlinearity for carcinogenicity. 

4.2.7	 Does the Use of a Maximum Tolerated Dose Affect Cancer Potency 

Estimates? 

The draft final cancer guidelines (USEPA, 2003c) include an extensive discussion of the 

role of dose selection in the conduct and interpretation of data from cancer tests carried out in 

animals. Section 2.2.2.1 deals with “Long-Term Carcinogenicity Studies,” and section 2.2.2.1.1, 
Page 62 Examination of EPA Risk Assessment Principles and Practices 

“Dosing Issues,” is in essence a discussion of the role of the MTD in cancer testing in animals 

and interpretation of data from such tests. 

Section 2.2.2.1.1 makes the following critical points, among others, to justify the use of 

MTDs: “...failure to reach a sufficient dose reduces the sensitivity of the studies.” Because 

animal tests are conducted with a relatively small number of animals, dosing at the MTD is 

necessary to provide the best chance that cancer will be seen in test animals if a chemical is 

carcinogenic. Further, the document recognizes that “...overt toxicity or inappropriate 

toxicokinetics due to excessively high doses may result in tumor effects that are secondary to the 

toxicity rather than directly attributable to the agent.” 

The document also stresses the importance of establishing that the MTD has not been 

exceeded. Criteria for determining this include whether “...an adequate high dose would 

generally be one that produces some toxic effects without unduly affecting mortality from effects 

other than cancer or producing significant adverse effects on the nutrition and health of the test 

animal.” Thus, our practice is that the experimental conditions are scrutinized to ensure that the 

MTD has been reached but not exceeded. We further believe that effects seen at the MTD may 

be appropriate for use in risk assessment when the data have been critically evaluated. 

4.2.8	 Are the Cancer Risks Estimated by EPA the Expected Incidence of Cancer 

for a Given Exposure? 

Our risk estimates are designed to ensure that risks are not underestimated, which means 

that a risk estimate is the upper bound on the estimated risk. In past guidelines (USEPA, 1986a), 

we have explicitly stated that the true cancer potency “could be as low as zero.” Since the CSF is 

generally multiplied by the exposure estimate to generate an estimated risk, a CSF of zero would 

result in a risk estimate of zero. 

If other procedures, e.g., odds ratios from epidemiologic studies, result in risk estimates 

that should be interpreted differently, these should be clearly delineated. In particular, if the 

method either requires or results in maximum likelihood estimates (MLEs) rather than bounding 

estimates being used, this should be clearly stated. 

According to Rhomberg (1997), OSHA uses the same methodology as EPA, but it 

presents and focuses on the MLE dose-response curve. CPSC also uses the same model as EPA, 

but modifies its constraints so that, with the limited data usually available, the MLE of the doseresponse function is linear at low doses. Rhomberg asserts that this procedure also results in an 

upper-bound estimate, although no proof is given in the cited document. 
Examination of EPA Risk Assessment Principles and Practices Page 63 

4.2.9 How Does EPA Evaluate Incidence Versus Mortality for Cancer Rates? 

Incidence of a disease is a measure of disease occurrence. Incidence rate measures the 

frequency of a given disease in a population. A concept closely related to incidence rate is that of 

the risk or probability of developing a disease over a period of time. The incidence rate for a 

disease is also referred to as the hazard rate for developing that disease (Kalbfleisch and Prentice, 

1980). Mortality rate, on the other hand, measures the frequency of death rate in a population 

from a given disease. Both incidence and mortality rates are measures of risk and have been used 

to calculate risk in different risk assessments. The appropriate measure depends upon the 

particular risk of interest and the type of data available. It is important that this be generally 

made clear in risk assessments. 

Incidence rates give a more accurate indication of risk of a disease in a population. 

However, obtaining information about incidence rates in a population is usually difficult. One 

place to look for incidence rates of site-specific cancers in a population is the Surveillance, 

Epidemiology, and End Results (SEER) data. Obtaining mortality data and computing rates is 

comparatively easy, because every death is recorded. Thus, over the years, the most commonly 

used data for risk assessments have been the mortality rates. An adjustment between mortality 

and incidence can be carried out using information from databases such as SEER. 

In diseases for which survival is poor and mortality is 95% to 100% within a short period 

of time (e.g., lung cancer, pancreatic cancer), the mortality rates are good surrogates for incidence 

rates. A risk assessment based on these mortality rates provides a good estimation of true risk in 

a given population. In diseases for which survival is higher and mortality is lower (e.g., nonmelanoma skin cancer, urinary bladder cancer), the mortality rates are poor surrogates for 

incidence rates. A risk assessment based on these mortality rates therefore provides an 

underestimation of risk in a given population.

 EPA’s draft final Guidelines for Carcinogen Risk Assessment (USEPA, 2003c) 

recognize this issue. Section 3.2.1 of the guidelines states that “Analysis of epidemiologic 

studies depends on the type of study and quality of data, particularly the availability of 

quantitative measures of exposure. The objective is a dose-response curve that estimates the 

incidence of cancer attributable to exposure to the agent.” It further notes that “The analysis is 

tailored to the nature of each study, with due consideration of the consequences of study design. 

For example, many studies collect information from death certificates, which leads to estimates 

of mortality rather than incidence. Because survival rates vary for different cancers, the analysis 

can be improved by adjusting mortality figures to reflect the relationship between incidence and 

mortality.” This was the approach used in EPA’s assessment of 1,3-butadiene, found in IRIS 

(USEPA, 2002e). A similar approach was used in the NRC’s assessment of the cancer risk from 

arsenic ingestion (NRC, 2001). 
Page 64 Examination of EPA Risk Assessment Principles and Practices 

It should be noted that when the CSF is derived from animal data, the extrapolation uses 

the tumor incidence data. 

4.2.10 If Data Are Only Available for One Route of Exposure, Does EPA Evaluate 

Other Routes of Exposure? 

Our general assumption is that chemicals that cause internal toxicity by one route of 

exposure will do so by any other route if that route also produces an internal dose. This position 

is based on the scientific principle that the internal dose to the tissue of interest is the ultimate 

determinant of toxicity. We advocate development and use of agent-specific pharmacokinetic 

information to inform this position. 

In its 1994 report, the NRC (NRC, 1994) states that “... the target-site dose is the ultimate 

determinant of risk ...” Thus, the manner or route of systemic exposure is secondary (except as it 

affects internal doses, e.g., by first-pass metabolism) to the internal dose. Others in government 

also reflect this fundamental principle (e.g., Corish and Fitzpatrick, 2002; ACGIH, 1991). 

Some chemicals have demonstrated the same (or very close to the same) target organs in 

response to chemical exposure, regardless of route. For example, for cumene, a kidney effect is 

observed after oral exposure, implying an internal dose of cumene or its metabolites to the kidney 

via the oral route; a kidney effect is also seen after an inhalation exposure to cumene, implying 

an internal dose of cumene or its metabolites to the kidney via the inhalation route. This 

demonstrates a target organ dosing independent of route of administration and a fundamental role 

for establishing an internal dose leading to toxicity. Three other examples from IRIS of chemical 

assessments having a common critical target organ via inhalation (RfC) and oral (RfD) routes of 

administration are shown below, in table 4-1. 

Table 4-1: Route Comparison 

Agent 

Target Organ Affected via the 

Oral Route Inhalation Route 

Cumene (CAS # 98-82-8) Kidney Kidney 

1,1-dichloroethylene (CAS # 75-35-4) Liver Liver 

Vinyl chloride (CAS # 75-01-4) Liver Liver 

Several examples from studies measuring kinetic endpoints of compounds administered 

by various routes of exposure found results to be mostly parallel across routes. This indicates 

similar kinetic processes regardless of route. 
Examination of EPA Risk Assessment Principles and Practices Page 65 

a) After exposing rats to benzo(a)pyrene via either oral, inhalation, or intravenous 

routes, Ramesh et al. (2002) observed data indicating metabolism to all principal 

metabolites across all routes. 

b) Kaneko et al. (1995) reported metabolism in rats of xylene by inhalation, oral, and 

intraperitoneal (ip) routes of exposure, with a significant “first-pass” effect via the 

oral route. This latter effect was observed only when the liver enzymes were 

altered by compounds known to induce activating enzymes. 

c) After administration of alkyl ketones to rats via oral or inhalation route, Duguay 

and Plaa (1995) observed dose-related increases in parent and metabolite in all 

tissues examined (plasma, liver, and lung). 

d) Gospe and Albayati (1994) demonstrated that oral administration of toluene 

produces blood toluene concentrations that can simulate blood levels achieved 

after inhalation exposures of this agent. 

e) Timchalk et al. (1991) observed that radiolabeled dichloropropane administered to 

rats was readily absorbed, similarly metabolized, and excreted after both oral and 

inhalation exposures. 

Studies measuring both toxicokinetics and toxicity administered by various routes of 

exposure found, in general, parallel results regardless of route. 

a)	 In a study measuring toxic in addition to kinetic endpoints, Gansewendt et al. 

(1991) reported that both oral and inhalation methyl bromide administration in 

rats resulted in the same levels and species of DNA adducts, with the highest 

levels found in the stomach and forestomach for both routes. 

b)	 After administering rats hepatotoxins that were either poorly (carbon 

tetrachloride) or highly (chloroform) metabolized, Wang et al. (1997) observed 

and reported that toxicodynamics were parallel and toxicity was present for both 

compounds via oral, inhalation, or ip routes. 

c)	 McEuen et al. (1995) noted that oral and ip routes of administration of 

dinitrobenzene produced significantly higher levels of active metabolite via the 

oral route but only subtle differences in target organ (testes) toxicity in rats. 

d)	 Using a PBPK model for cumene to examine measures of dose associated with 

renal toxicity across an oral and an inhalation study in rats, Foureman and Clewell 
Page 66 Examination of EPA Risk Assessment Principles and Practices 

(1999) noted that measures of dose generated from either study were associated 

with the toxicity in a dose-related manner, regardless of route. 

What Data May Set Aside the Route-to-Route Extrapolation? 

The draft final cancer guidelines (USEPA, 2003c) discuss some evidence that could 

modify the application of this extrapolation, stating that an exception would occur when there is 

convincing toxicokinetic evidence that absorption for a given chemical does not occur by another 

route of interest. The basis for this exception is establishing the absence of an internal dose such 

that the extrapolation may not be applied and such a chemical could be characterized as not being 

toxic via this particular route of administration. 

In the absence of data to the contrary, the route-to-route extrapolation assumption used is 

that if a chemical is absorbed via one route, it will also be absorbed by all other routes. 

Demonstration of any degree of uptake for each of the routes of interest is sufficient to allow the 

qualitative judgment to apply the route-to-route extrapolation. The level of information 

necessary to set aside the extrapolation (i.e., “contrary” evidence ) may largely be a function of 

the individual stressor being tested. The similarities of processes and structures involved in 

absorptive processes across biological systems may be regarded as an unstated principle that 

underlies the extrapolation. Absorption obligates passing through a tissue and, fundamentally, 

tissues have more similarities than differences. The presumption is that if the stressor can 

penetrate tissue via one route, it can penetrate the tissue if it arrives there via another route. Thus 

it is anticipated that most agents are absorbed to some degree into biological systems regardless 

of route (e.g., see results summarized in Owen, 1990). 

Chemicals causing point-of-entry toxicity are often very reactive, and thus are less likely 

than other chemicals to produce an internal dose. Chemicals that do not establish an internal 

dose by a given route would not be presumed to give internal toxicity by that route. Evidence of 

robust point-of-entry toxicity by an agent may thus be considered contrary evidence for the 

contention that toxicity by one route of exposure by that agent is presumed to do so by another 

route. IRIS lists assessments for a number of highly reactive agents that elicit only clear and 

robust portal-of-entry effects — in the context of this issue, contrary evidence. Nearly all of 

these assessments explicitly consider that these agents would not be anticipated to elicit internal 

toxicity and discount the need for information on systemic toxicity. This discounting is done by 

decreasing or even eliminating the “database” UF typically evoked at a value of 10 when studies 

such as developmental and reproductive toxicity (toxicities typically associated with an internal 

dose) are absent. Table 4-2 gives some examples. The practice of omitting the database UF 

when portal-of-entry effects occur may not be followed in some older IRIS assessments that have 

not yet been updated. For example, a database UF of 3 is applied specifically for missing 

developmental data in the assessment of one other of the isocyanates, TDI (CAS # 26471-62-5). 
Examination of EPA Risk Assessment Principles and Practices Page 67 

A database UF of 10 is also applied in the RfC for 2-chloroacetophenone (CAS # 532-27-4), a 

highly reactive tearing agent. 

Table 4-2: Point-of-Entry Toxicity Examples 

Agent (CAS#) Valuea and Commentary of “Database” UF (Online IRIS) 

Acrolein (107-02-8) UF = 1, not applied due to lack of systemic distribution (2002) 

Ammonia (7664-41-7) UF = not greater than 3, no significant distribution likely to occur 

(1991) 

Epoxybutane 

(106-88-7) 

UF = 1, extrarespiratory circulation of EBU minimal (1992) 

HCl (7647-01-0) UF = 1, expected portal-of-entry effect, UF not applied (1995) 

HDI (822-06-0) UF = 3, unlikely that HDI is distributed in significant amounts 

(1994) 

H3PO4 (7664-38-2) UF = 1, no UF to be applied (1995) 

MDI (101-68-8) UF = 3 (1998) 

Methacrylate (80-62-6) UF = 1, reactive, potential for systemic effects is remote (1998) 

a

 This UF has a maximum value of 10. 

4.3 Use of Uncertainty Factors 

4.3.1 How Are NOAELs or LOAELs Used for Extrapolations? 

When developing a non-cancer reference value (i.e., a RfD or RfC) for a chemical 

substance, EPA surveys the scientific literature and selects a critical study and a critical effect to 

serve as the point of departure for the assessment. The critical effect is defined as the first 

adverse effect, or its known precursor, that occurs in the most sensitive species as the dose rate of 

an agent increases (USEPA, 2002c). Such a study, whether an occupational human study, a 

deliberately dosed animal study, or some other study, typically involves exposure at a range of 

doses. The highest 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 

appropriate control is — as mentioned above — called the “no-observed-adverse-effect level” 

(NOAEL). When a NOAEL can be identified in a critical study, it becomes the basis of the 

reference value derivation. The NOAEL is divided by appropriate UFs (e.g., for intraspecies 
Page 68 Examination of EPA Risk Assessment Principles and Practices 

variation or study duration) to derive the final reference value. If NOAEL cannot be identified, 

then a “lowest-observed-adverse-effect level” (LOAEL) — also mentioned above — is identified 

instead. A LOAEL is the lowest exposure level at which there are biologically significant 

increases (with or without statistical significance) in frequency or severity of adverse effects 

between the exposed population and its appropriate control group. The NOAEL is generally 

presumed to lie between zero and the LOAEL, so an UF (generally 10 but sometimes 3 or 1) is 

applied to the LOAEL to derive a nominal NOAEL. Other factors are then applied to derive the 

reference value. More recently, some assessments have used a BMD approach instead of the 

traditional NOAEL/LOAEL approach (USEPA, 2000g). 

For each non-cancer reference value (i.e., RfC or RfD) for a chemical substance in IRIS, 

the following table shows whether that value was based on a NOAEL, a LOAEL, or a benchmark 

dose lower confidence level. The table clearly shows that the majority of non-cancer reference 

values in IRIS are based on NOAELs. 

Table 4-3: Effect Levels Used To Derive Non-Cancer Reference Values in IRIS 

RfCs RfDs 

NOAEL 35 294 

LOAEL 21 51 

BMD 13 10 

Total 69 355 

4.3.2	 What Uncertainty Factors Does EPA Use To Reduce the Experimental 

NOAEL in Health Assessments? 

The Risk Assessment Forum RfD/RfC Technical Panel report (USEPA, 2002c) defines 

an uncertainty factor as: 

one of several, generally 10-fold, default factors used in operationally deriving the RfD and the RfC from 

animal experimental data. The factors are intended to account for (1) variation in sensitivity among the 

members of the human pop ulation; (2) the uncertainty in extrapolating animal data to humans; (3) the 

uncertainty in extrapolating from data obtained in a study with less-than-lifetime exposure to lifetime 

exposure; (4) the uncertainty in extrapolating from a LO AEL rather than from a NO AEL; and (5) the 

uncertainty associated with extrapolation when the database is incomplete. The exact value of the UFs 

cho sen d epe nd o n the q uality of the stu dies av ailable , the exten t of the d ataba se, and scientific ju dgm ent. 

Some investigators evaluated the accuracy and limitations of allocating a value of 10 for 

each area of uncertainty (Dourson and Stara, 1983; Woutersen et al., 1984; Kadry et al., 1995; 
Examination of EPA Risk Assessment Principles and Practices Page 69 

Dourson et al., 1996). For example, the subchronic-to-chronic UF, some conclude, is probably 

closer to a two- to threefold difference of uncertainty (95% of the time) and a UF of 10 should be 

considered as a loose upper-bound estimate of the overall uncertainty. Others (see Dourson and 

Stara, 1983, for survey of arguments representing various points of view) conclude that a 10-fold 

interspecies UF could account for all animal to human differences and likewise for interhuman 

variability. There are opposing arguments that EPA’s application of UFs may not be 

conservative enough (Dourson and Stara, 1983). 

EPA applies UFs in health assessments based on the available data and the scientific 

judgment of EPA risk assessors and internal and external peer reviewers. In cases where 

chemical-specific data are lacking, each UF is typically no greater than 10. For example, the 

majority of IRIS Toxicological Reviews provide justifications for the individual UFs selected for 

a particular chemical substance. The choices of critical health endpoint, principal study, and UFs 

undergo rigorous internal and independent, external scientific peer review before being presented 

in the IRIS database. 

For several years, EPA has used a more qualitative approach to modify the usual 10-fold 

default values. For example, in deriving inhalation RfC values, the interspecies variability UF of 

10 is used in the absence of data, where the distribution is assumed to be log-normally 

distributed. While EPA has not yet established guidance for the use of chemical-specific data for 

derived UFs, the reference concentration methodology guidance (USEPA, 1994a) provides 

opportunities for using data-derived interspecies UFs by subdividing the factor of 10 to allow for 

separate evaluations of toxicokinetics and toxicodynamics. The advantage to such subdivision is 

a default UF of 10 for interspecies variability that can now be reduced to 3 when animal data are 

dosimetrically adjusted to account for toxicokinetics (USEPA, 2002c). 

4.3.3 Does EPA Consider the Effects of Combining Several UFs? 

In its review of the RfD/RfC practice in the Agency, the RfD/RfC Technical Panel 

reviewed the application of UFs in health assessments (USEPA, 2002c). The Panel recognized 

the potential for overlap in the individual UFs and concluded that the application of five UFs of 

10 for the RfD/RfC is inappropriate. Therefore, due the uncertainty inherent in values when so 

many UFs are applied, the Panel recommended that no reference value (RfD/RfC) for any 

particular chemical substance be derived if the composite UF is greater than 3,000. It further 

recommended avoiding the derivation of a reference value that involves application of the full 

10-fold in four or more areas of extrapolation. The report also recommended to discontinue the 

use of the modifying factor as a UF. 
Page 70 Examination of EPA Risk Assessment Principles and Practices 

The two graphs below show EPA’s historical practice for combining UFs to derive a RfD 

or RfC. As can be seen, when reference values were originally being generated, some 

combinations of UFs reached as high as 10,000. As we gained greater and greater experience 

and obtained more usable data, the trend has shown a decrease in the combination of UFs to such 

high levels. The more recent RfDs and RfCs are more in line with the recommendation from the 

Technical Panel: no derived reference value exceeds 3,000. 

Figure 4-1: Use of Uncertainty Factors Over Time (to Set RfDs and RfCs) 
Examination of EPA Risk Assessment Principles and Practices Page 71 

4.3.4 Does EPA Use UFs Derived for Specific Chemicals? 

With respect to departing from the default UFs with data-derived or chemical-specific 

adjustment factors, the RfD/RfC Technical Panel indicated that guidance on how to use 

chemical-specific adjustment factors in risk assessment will be developed. Toward that end, 

EPA has recently begun to consider data-derived approaches for chemical-specific adjustment 

factors in risk assessments. 

4.4 Weight of Evidence 

4.4.1 How Does EPA Consider Weight of Evidence? 

Risk assessment involves consideration of the weight of evidence provided by all 

available scientific data. In other words, “weight of evidence evaluation is a collective 

evaluation of all pertinent information so that the full impact of biological plausibility and 

coherence is adequately considered” (USEPA, 1999b). Judgment on the weight of evidence 

involves consideration of the quality and adequacy of data and consistency of responses induced 

by the stressor. The weight-of-evidence judgment requires combined input of relevant 

disciplines: toxicology, biology, chemistry, epidemiology, statistics, etc. Initial views of the 

database may change significantly when other data are brought into consideration. For example, 
Page 72 Examination of EPA Risk Assessment Principles and Practices 

the impact of a positive animal carcinogenicity study may be diminished by high-quality negative 

studies; or a weak association in human epidemiologic studies may be bolstered by consideration 

of other key data from animal or other assays. Generally, no single study, whether positive or 

negative, drives the overall weight-of-evidence judgment. And study findings are not scored by 

any mathematical algorithm; rather, they are based on professional scientific judgment. 

4.4.2	 Does EPA Give More Weight to Positive Animal Studies Than to Negative 

Animal Studies? 

EPA evaluates all data, including both “positive” and “negative” studies, in assessing the 

weight of evidence for toxicity of a chemical. Assessment decisions are, by nature, subject to 

professional judgment and require input from experts in a variety of disciplines including 

toxicology, pathology, and statistics. Generally, EPA does consider positive animal studies even 

in the presence of negative studies. Negative studies do not “outweigh” positive studies, though, 

unless they present a robust, compelling outcome relative to the positive studies. 

To examine the comment that EPA always gives more weight to positive animal studies 

than to negative studies, 108 chemicals classified as Class D (not classifiable as to human 

carcinogenicity) in IRIS were searched for evidence as to whether they included any positive 

animal studies. Seventeen of those substances were found to have statistically significant 

positive animal data, but it was the professional judgment of those performing the assessment 

and those who scientifically peer reviewed it that these substances were indeed nonclassifiable as 

to human carcinogenicity. Thus, we do not automatically translate positive animal 

carcinogenicity data into classification as a human carcinogen. Rather, it is EPA’s practice to 

base decisions on the best available science. 

4.4.3	 Does EPA Sometimes Weight Animal Data More Than Human Data? 

EPA prefers high-quality human studies over animal studies because they provide the 

most relevant kind of information for human hazard identification. In the absence of usable 

human data, the default assumption is that positive effects in animal cancer studies indicate 

carcinogenic potential in humans. Risk analysis takes all studies into account, whether they show 

positive associations, null results, or even protective effects. In weighing positive versus null 

studies, those judged to be of high quality are given more weight than those judged 

methodologically less sound, and possible reasons for inconsistent results should be sought. Null 

results from a single epidemiologic study cannot prove the absence of health effects because of 

the limitations that may be associated with the study — inadequate statistical power, inadequate 

design, imprecise estimates, confounding factors, and others. However, null results from a welldesigned and well-conducted epidemiologic study that contains usable exposure data can help 

define upper limits for the estimated dose of concern for human exposure if the overall weight of 

evidence indicates the agent is a potential human carcinogen. 
Examination of EPA Risk Assessment Principles and Practices Page 73 

EPA tries to use human studies whenever feasible and appropriate. As the NRC pointed 

out (NRC, 1994), “Epidemiologic studies clearly provide the most relevant kind of information 

for hazard identification, simply because they involve observations of human beings, not 

laboratory animals.” As a result, human data form the basis of a number of highly visible 

assessments such as those of arsenic, chromium, nickel, benzene, and vinyl chloride. It is true, 

however, that we have relied more heavily on animal data for some assessments even when 

human data were available. The case of epichlorohydrin is an example of such an instance. 

Despite the fact that two retrospective cohort mortality studies of exposed workers found no 

evidence of an increased cancer risk, epichlorohydrin is classified as a “probable human 

carcinogen.” The apparent discounting of the human data is explained as follows: 

The first study is inadequate for valid carcinogenicity assessment because of low exposure, short exposure 

duration, short study period, and the young age of the cohort. The second study suffers from some of the 

same limitations, as well as that of a small cohort size with at least 10 years exposure (274 individuals). 

While pointing out the benefits of epidemiologic evidence, the NRC noted that the 

obvious and substantial advantage (of human studies) is offset to various degrees by the 

difficulties associated with obtaining and interpreting epidemiologic information (NRC, 1994). 

As the Office of Science and Technology Policy indicated (OSTP, 1985), the epidemiologic 

method (in the assessment of cancer risk) is often hampered by the long latent period between 

exposure to a carcinogenic agent and the development of cancer, by the inability to control for 

the confounding influences of unknown risk factors, by problems in assessing specific agents 

when the human exposures are to mixtures, by the frequent absence of appropriate groups for 

study, and by a variety of difficulties associated with accurate and unbiased historical exposure 

assessment or disease ascertainment. In short, the Agency, while recognizing the clear 

advantages of epidemiologic studies, also recognizes the limitations of these studies; and it is 

these limitations which sometimes constrain the use of human data in an assessment. 

For some comparisons of the use of animal versus human studies in risk assessment 

decisions within the IRIS database, please see the tables below. 

Animal Versus Human Studies Tables 

All substances in the IRIS database were searched for the keyword “human,” which 

provided 541 matches (essentially the whole database). From these matches, the first 60 were 

examined to identify compounds for which human data are available, yet animal studies serve as 

the basis for assigning RfCs, RfDs, or cancer risk estimates. 

Within the first 60, there were 15 with RfC listings. Of those 15, 8 list specific human 

data in the inhalation exposure section. Four of those eight use human data as the principal study 

guiding determination of the RfC, and four did not. Table 4-4 outlines the explanations provided 

for why the available human data were not a principal study for the latter four substances. The 
Page 74 Examination of EPA Risk Assessment Principles and Practices 

table also presents brief statements of positive and negative results from listed human studies and 

interspecies UFs. 

Among the first 60, there were 42 with RfD listings. Of these 42, nine list specific human 

data in the oral exposure section. Five use the human data as the principal study in guiding 

determination of the RfD; four do not. Table 4-5 outlines the explanations provided for these 

four studies as to why they were not used to set the RfD. 

There were 27 among the first 60 with classifications of carcinogenicity. Of these 27, 10 

identify human data in the assessment. Four of these assessments use human data to make the 

classification of carcinogenicity; six do not. Table 4-6 provides the explanations as to why, for 

the latter six substances, human data were not used to make the carcinogenic classification. 

Table 4-4: Explanations of Why Human Results Were Not Used To Estimate the RfC for Four 

Substances in the IRIS Databasea 

Chemical Human 

Data (HD) 

Listed 

HD = 

Principal 

Study 

Explanation 

Bromomethane Yes No Several studies have been conducted on the longer-term 

effects o f occu patio nal exp osure bro mom ethane . None 

of these studies can serve as the basis for the 

derivation of an RfC for bromomethane because of 

concurren t exposures to other chem icals, inadequ ate 

quantitation of exposure levels and/or durations, and 

other deficits in study design. 

[UF ] of 3 for interspecies ex trapolation b ecause 

dosimetric adjustments have been applied. 

• Positive human occupational study: neurological 

effects (A nge r et al., 19 86). 

• Fo ur ca ses of re por ted o ccu patio nal to xicity (H erzstein 

and Cu llen, 19 90). 
Examination of EPA Risk Assessment Principles and Practices Page 75 

Dichlorvos Yes No Justification not provided. A UF of 3 is used for 

interspecies extrapolation. 

• Limited reporting of mixed results in study of 

reside ntial fu mig ators and reside nts (G old e t al., 

198 4). 

• Plasm a cho linesterase activity an d vitam in A de crease 

in five of six pesticide m anu facturing p lant wo rkers 

(Em ber e t al., 19 72). 

• Ca se stud ies of p oison ings re por t clinica l and selfrepo rted sy mp tom s (Low et al., 19 80; Ree ves et a l., 

198 1). 

• Decrease in plasma cholinesterase activity, but no 

significant difference in reported clinical symptoms 

between test and control subjects in volunteer human 

studie s (Ride r et al., 19 67; Slom ka a nd H ine, 1 981 ). 
Page 76 Examination of EPA Risk Assessment Principles and Practices 

Table 4-4: Explanations of Why Human Results Were Not Used To Estimate the RfC for Four 

Substances in the IRIS Databasea 

Chemical Human 

Data (HD) 

Listed 

HD = 

Principal 

Study 

Explanation 

Ac rylo nitrile Yes No Justification not provided. 

An uncertainty factor of 3 for interspecies variability is 

used because the use of the dosimetric adjustments 

account for part of this area of uncertainty. 

• No excess mortality for non-cancer endpoints in human 

occ upa tiona l epi stud ies (O’ Ber g, 19 80; Ott et a l., 

1980; Werner and Carter, 1981; O’Berg et al., 1985; 

Ch en et a l., 198 7). 

• Self-re por ted a nd c linical sy mp tom s app aren t in 

occ upa tiona l study (Wilson et al., 19 48). 

• No effects in h um an v olun teer stud y (Jak ubo wski e t al., 

198 7). 

Propylene glycol 

monomethyl ether 

(PGME) 

Yes No Comment on use of human data: A human exposure study 

(Stewart et al., 1970) indicates that the RfC should 

prevent irritation. Additional justification not provided. 

[An uncertainty] factor of 3 is used for interspecies 

extrapolation given the dosimetric adjustment and the 

zero-order pharmacokinetic rate of elimination of PGME 

that is suggestive of an adaptive metabolic response 

(Morgott and Nolan, 1987). 

• Hum an volunteer study was positive for self-reported 

symptoms but negative for clinical effects (Stewart et 

al., 19 70). 

a

 These four substances were identified from the first 60 matches in the IRIS data file for which the word “human” 

was identified. RfCs were estimated for 15 of these 60 substances. Human data relevant to the RfC were 

available for 8 of the 15 substances. The human data were used as the principal study for the RfC for four of 

those eight substances. This table describes the reasoning provided as to why the human data were not used in the 

other four instances. Interspecies UFs are listed. Results of human studies are included in italics. 
Examination of EPA Risk Assessment Principles and Practices Page 77 

Table 4-5: Explanations of Why Human Results Were Not Used To Estimate the RfD for Four 

Substances in the IRIS Databasea 

Chemical Human 

Data (HD) 

Listed 

HD = 

Principal 

Study 

Explanation 

Chromium 

(VI) 

Yes No In 1965, a study of 155 subjects exposed to drinking water at 

concentrations of approximately 20 mg/L was conducted outside 

Jinzhou, China ... Precise exposure concentrations, exposure 

durations, and confounding factors were not discussed, and 

this study does not prov ide a NO AEL for the observed effects. 

However, the study suggests that gastrointestinal effects may occur 

in humans following exposures to hexavalent chromium at levels 

of 20 ppm in drinking water (Zhang and Li, 1987). 

[UF] — two 10-fold decreases in dose to account for both the 

expected interhuman and interspecies variability in the toxicity of 

the chemical in lieu of specific data. 

• Po sitive drin king wa ter stud y (Zh ang and Li, 19 87). 

• Studies of positive allergic response (Bruynzeel et al., 1988; 

Po lak, 1 983 ; Cro nin, 1 980 ; Hu nter, 1 974 ). 
Page 78 Examination of EPA Risk Assessment Principles and Practices 

Table 4-5: Explanations of Why Human Results Were Not Used To Estimate the RfD for Four 

Substances in the IRIS Databasea 

Chemical Human 

Data (HD) 

Listed 

HD = 

Principal 

Study 

Explanation 

Uranium, 

soluble salts 

Yes No Uranium is a classical nephrotoxic. The toxicity of this chemical 

to humans has been of interest since the 1800’s when uranium was 

used as a homeopathic cure for diabetes mellitus (Hodge, 1973). 

These early reports demonstrate the susceptibility of humans to the 

nephrotoxicity of ingested uranium, but pro vide inad equ ate b asis 

for estima ting the th reshold dose fo r toxic effec ts. 

[UF of] 10 for both intraspecies and interspecies variability to the 

toxicity of the chemical in lieu of specific data. 

• Rep orts of n eph rotox icity with thera peu tic use (H odg e, 19 73). 

• Mixed results from experiments on uranium excretion and 

toxicity (Hursh and Spo or, 197 3; Lusse nho p et al., 195 8; Hu rsh 

et al., 19 69). 

Naphthalene Yes No Humans exposed via inhalation, combined inhalation and dermal 

exposure, and combined inhalation and oral exposure have 

developed hemolytic anemia. Hemolytic anemia is characterized 

by findings of lowered hemoglobin, hematocrit, and erythrocyte 

values, elevated reticulocyte counts, Heinz bodies, elevated serum 

bilirubin, and fragmentation of erythrocytes. In severe cases, the 

hemolytic anemia was accompanied by jaundice, high serum levels 

of bilirubin, cyanosis, and kernicterus with pronounced 

neurological signs. Ne ither ora l nor inha lation expo sure levels 

w ere a vaila ble in h uma n stu dies re por ting a nem ia (MelzerLange and W alsh-Kelly, 1989; Owa, 1989; Owa et al., 1993). 

[UF of] 10 to extrapolate from rats to humans. 

• Human studies positive for hemolytic anemia (Melzer-Lange 

and Wa lsh-K elly, 19 89; Ow a, 19 89; Ow a et a l., 199 3). 
Examination of EPA Risk Assessment Principles and Practices Page 79 

Table 4-5: Explanations of Why Human Results Were Not Used To Estimate the RfD for Four 

Substances in the IRIS Databasea 

Chemical Human 

Data (HD) 

Listed 

HD = 

Principal 

Study 

Explanation 

1,3-

Dichloropropene 

Yes No There are no ch ronic h uma n studies suita ble for d ose-resp onse 

assessment. Chronic feeding studies in rats (Stott et al., 1995) 

and mice (Redmond et al., 1995) and chronic gavage studies 

(NTP, 1985) using both species are available. The feeding studies 

are favored over the gavage studies because the route of 

administration is more relevant to human exposure. 

The toxic okin etics in h uma ns ar e simila r to th ose o bser ved in 

rats. Waechter et al. (1992) showed that the absorption of 1,3-

dichloropropene from inhalation exposure of humans (72%–82%) 

was similar to absorption in rats (82%; Stott and Kastl, 1986). 

The default uncertainty factor of 10 for interspecies extrapolation 

is applied because there are no data on the relative sensitivity of 

rats and humans to stomach irritation. 

• Hu ma n stud ies dem ons trate co ntac t derm atitis (Bo usem a et a l., 

199 1; N ater a nd G oos ken s, 197 6). 

• Hum an poisonings result in neurotoxic symptoms (Hernandez et 

al., 19 94). 

a

 These four substances were identified from the first 60 matches in the IRIS data file for which the word “human” 

was identified. RfDs were estimated for 42 of the 60 substances. Human data relevant to the RfD were available 

for 9 of the 42 substances. The human data were used as the principal study for the RfC for five of the nine 

substances. This table describes the reasoning provided as to why the human study were not used in the other 

four instances. Interspecies uncertainty factors are listed. Results of human studies are included in italics. 
Page 80 Examination of EPA Risk Assessment Principles and Practices 

Table 4-6: Explanations of Why Human Results Were Not Used for the Carcinogenic 

Classification for Six Substances in the IRIS Databasea 

Chemical Human Data 

(HD) Listed 

Explanation 

Bromomethane Yes D (N ot classifiable as to human carcinogenicity) 

Inadequate human data. A prospective mortality study was reported for 

a population of 3579 white male chemical workers. The authors noted 

that it was difficult to draw definitive conclusions as to causality 

because of the lack of exposure information and the likelihood that 

exposure was to many brominated compounds. 

• (Wong et al., 1984) study looked at cancer mortality in the above 

mentioned occupational cohort. Only increased cancer mortality was 

from testicular c ance r. 

p,p’-

Dic hlor odip hen yltrichloroethane 

(DDT) 

Yes B2 (Probable human carcinogen — based on sufficient evidence of 

carcinogenicity in animals) 

Inadequate human data. The existing epidemiological data are 

inadequate. 

• Autopsy studies relating tissue levels of DDT to cancer incidence 

have yielded conflicting results. Three studies reported that tissue 

levels of DD T and DD E we re highe r in canc er victims than in those 

dying of other disea ses (Casare tt et al., 1968; D acre an d Jenn ings, 

197 0; W asser ma n et a l., 197 6). In o ther stu dies n o suc h rela tionsh ip 

wa s seen (Ma ier-B ode , 196 0; R obin son et al., 19 65; Ho ffma n et a l., 

196 7). Studies of o ccup ationa lly exposed worke rs and v oluntee rs 

have been of insufficient duration to be useful in assessment of the 

carcino genicity of D DT to hum ans. 
Examination of EPA Risk Assessment Principles and Practices Page 81 

Table 4-6: Explanations of Why Human Results Were Not Used for the Carcinogenic 

Classification for Six Substances in the IRIS Databasea 

Chemical Human Data 

(HD) Listed 

Explanation 

Polychlorinated 

biphen yls (PC Bs) 

Yes B2 (Probable human carcinogen — based on sufficient evidence of 

carcinogenicity in animals) 

Inadequate human data. The human studies are being updated; currently 

available evidence is inadequate, but suggestive. 

• Sign ificant c anc er exc esses in hum an c oho rt studie s (Berta zzi et al., 

198 7; B row n, 19 87; Sink s et al., 19 92). 

• Inconclusive results of occupational studies due to design limitations 

(NIO SH , 197 7; G ustav sson et al., 19 86; Sha lat et al., 1 989 ). 

• Liver cancer and skin disorders in populations exposed to PCBcontaminated rice oil in Japan and Taiwan (ATSDR , 1993; Safe, 

199 4). 

Die ldrin Yes B2 (Probable human carcinogen — based on sufficient evidence of 

carcinogenicity in animals) 

Inadequate human data. Two studies of workers exposed to aldrin and 

to dieldrin reported no increased incidence of cancer. Both studies were 

limited in their ability to detect an excess of cancer deaths ... Exposure 

was not quantified, and workers were also exposed to other 

organochlorine pesticides (endrin and telodrin). The number of workers 

studied was small, the mean age of the cohort (47.7 years) was young, 

the number of expected deaths was not calculated, and the duration of 

exposure and of latency was relatively short. 

• No statistically significant excess in cancer deaths among 

occupational cohort (Van Raalte, 1977; Ditraglia et al., 1981). 

~ same studies as cited in the aldrin assessment 
Page 82 Examination of EPA Risk Assessment Principles and Practices 

Table 4-6: Explanations of Why Human Results Were Not Used for the Carcinogenic 

Classification for Six Substances in the IRIS Databasea 

Chemical Human Data 

(HD) Listed 

Explanation 

Acrylamide Yes B2 (Probable human carcinogen — based on sufficient evidence of 

carcinogenicity in animals) 

Inadequate human data. There are two studies on the relationship of 

workers exposed to acrylamide and cancer mortality. A basic limitation 

of both studies is that the design is insufficient to derive an inference of 

relative risk. 

Based on inadequate human data and sufficient evidence of 

carcinogenicity in animals; significantly increased incidences of benign 

and/or malignant tumors at multiple sites in both sexes of rats, and 

carcinogenic effects in a series of one-year limited bioassays in mice by 

several routes of exposures. The classification is supported by positive 

genotoxicity data, adduct formation activity, and structure-activity 

relationships to vinyl carbamate and acrylonitrile. 

• No statistically significant increase in cancer mortality in human 

study (Collin s, 198 4). 

• Co nfou nde d resu lts in hu ma n stud y (So bel et a l., 198 6). 

Ald rin Yes B2 (Probable human carcinogen — based on sufficient evidence of 

carcinogenicity in animals) 

Inadequate human data. Two studies of workers exposed to aldrin and 

dieldrin (a metabolite of aldrin) did not find these workers to have an 

excess risk of cancer. B oth studies, however, were limited in their 

ability to detect an excess of deaths from cancer ... Exposure was not 

quantified, and workers were also exposed to other organochlorine 

pesticides (endrin and telodrin). A small number of workers was 

studied, the mean age of the cohort (47.7 years) was low, the number of 

expected deaths was not calculated, and the duration of exposure and of 

latency was relatively short. 

• No statistically significant excess in cancer deaths among 

occ upa tiona l coh ort (Va n R aalte , 197 7; D itraglia et al., 19 81). 

~ same studies as cited in the dieldrin assessment 

a

 These six substances were identified from the first 60 matches in the IRIS data file in which the word “human” 

was identified. O f these 60, 27 were classified as to their carcinogenicity. Of the 27, 10 identify human data in 

the assessment, and 4 use human data for identifying a carcinogenic classification. This table describes the 

reasoning provided as to why the human data were not used in the other six instances. Results of human studies 

are included in italics. 
Examination of EPA Risk Assessment Principles and Practices Page 83 

4.5 Toxicity Equivalency Factor Approach 

Several comments stated that the use of Toxicity Equivalency Factors (TEFs) is 

inappropriate because TEFs are based on the unproven assumption of additive toxic effects, 

compounded by the uncertainties inherent in the generation of TEFs, their application, and the 

heterogeneity of effects for which they are being used. However, some comments acknowledge 

the value of TEFs in certain circumstances. 

4.5.1 What Are TEFs? 

In 1986, EPA published the Guidelines for the Health Risk Assessment of Chemical 

Mixtures (USEPA, 1986b). These provide the bases for evaluating human risk from exposure to 

combinations of two or more chemical substances, regardless of source or of spatial or temporal 

proximity. They were extensively peer reviewed, submitted for public comments, and reviewed 

again by EPA’s Science Advisory Board. The guidelines have been supplemented by a technical 

support document (USEPA, 1988a) and the Supplementary Guidance for Conducting Health 

Risk Assessment of Chemical Mixtures (USEPA, 2000h). These guidance documents refer the 

assessor to several different methods for evaluating mixtures data of varying types: 

epidemiologic or toxicological data on the environmental mixture (the whole mixture) to which 

exposures occur; test data on a tested mixture judged to be sufficiently similar to the 

environmental mixture; data from a group of similar mixtures; and data on the mixture 

components. TEFs are used in evaluating mixture toxicity from data on components. Their use 

is based on an assumption of additivity among similarly acting components of a mixture. 

In the supplementary mixtures guidance (USEPA, 2000h), the Agency provided 

procedures for developing the relative potency factor (RPF) method and described TEFs as a 

special case of RPFs. The RPF method is component-based and relies on two types of 

information: toxicological dose-response data for at least one component of the mixture being 

addressed (referred to as the index compound or IC) and scientific judgment as to the toxicity of 

the other individual compounds in the mixture and of the mixture as a whole. If data are limited, 

the applicability of a set of RPFs may be restricted to certain effects, a specific route of exposure, 

or exposure duration. Application may also be limited to a certain portion of the dose-response 

curve. TEFs are a special case of RPFs, because they can be applied to any effect, route, or 

exposure duration, and across the entire dose-response curve (USEPA, 2000h). They are thus 

used only in relatively data-rich situations. Both RPFs and the special case of TEFs are intended 

to serve as interim approaches for addressing any mixture, pending the development of new data 

on the mixture’s toxicity. The supplementary mixtures guidance (USEPA, 2000h) noted an early 

Agency application of an RPF-like approach for polycyclic aromatic hydrocarbons (PAHs) 

(USEPA, 1993a; Schoeny et al., 1998). EPA has since applied the RPF method to 

organophosphate pesticides in the Guidance on Cumulative Risk Assessment of Pesticide 
Page 84 Examination of EPA Risk Assessment Principles and Practices 

Chemicals That Have a Common Mechanism of Toxicity (USEPA, 2002f). This was followed 

with the publication of biostatistical criteria for developing RPFs (USEPA, 2003e). 

The RPF method (including TEFs) is based on dose addition and assumes that the 

chemicals in a mixture share a common toxic mode of action that is relevant to the health 

endpoints being assessed (see section 5.7.2 for a complete discussion of dose addition). 

Operationally, this means that mixture components tested in the same bioassay should have doseresponse curves of similar shape between the toxicity thresholds and the maximum response. 

The components are assumed to be true toxicological representations of each other, although 

their relative toxic potencies may differ. 

When one uses the RPF method, it is necessary to identify the constraints of its 

application. For example, a set of RPFs may be restricted to oral exposures and not be usable for 

exposures to the same mixture through the inhalation route; this was the recommendation for 

RPFs applied to PAHs. TEFs for chlorinated dibenzo-p-dioxins, dibenzo-p-furans, and coplanar 

polychlorinated biphenyls (PCBs) have no identified constraints. 

To apply the method (see the following formula), one estimates an RPF for each mixture 

component; that is, one estimates component toxicity relative to that of the IC. 

Equation 4.5-1 

Where: 

Rm = risk posed by chemical mixture 

f1 (*) = dose-response function of index chemical 

Di = dose of the ith mixture component (i = 1, ..., n) 

RPFi = toxicity proportionality constant relative to index chemical for the i

th mixture 

component (i = 1, ..., n) 

n = number of chemical components in the mixture 

One estimates the RPF by choosing a response level and calculating the ratio of the dose causing 

that response for the component and the IC. For example, say that the ED10 (the effective dose at 

which 10% of a test population exhibits an effect) for component A is 500 micrograms (μg) and 

that of the index chemical is 5 μg. If the index chemical RPF is set at 1, the component RPF 

would be 5/500 or 0.01. The next step is to calculate the IC equivalent dose for the mixture. The 

equivalent dose of an individual component is the product of the amount of that component in 

the mixture and the RPF of the component. In our example, if there is 100 μg of component A in 

the mixture, then its IC equivalent dose is 100 μg × 0.01, or 1 μg. These equivalent doses are 

summed across all components to determine the mixture total equivalent dose. Lastly, one 
Examination of EPA Risk Assessment Principles and Practices Page 85 

estimates the risk posed by the mixture by comparing its summed IC equivalent dose to the doseresponse function (potency) of the IC. 

TEFs have been developed for three structurally related groups of compounds which have 

been demonstrated to exert similar biochemical and toxic endpoints: the polychlorinated 

dibenzo-p-dioxins (PCDDs), polychlorinated dibenzofurans (PCDFs), and (co-planar) PCBs. Of 

these, 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) is the most extensively researched and has 

served as the IC. Many studies support the TEF approach for these compounds because they 

share a common mechanism of action involving binding to a cellular receptor (AhR). TEFs were 

derived for 29 dioxin-like agents, which express each congener’s toxicity as a fraction of that of 

TCDD, the IC (van den Berg et al., 1998). 

4.5.2 How Did the TEF Approach Evolve? 

In the 1970s and 1980s, human health risk assessments of complex mixtures of PCDDs 

and PCDFs were generally done only by considering 2,3,7,8-TCDD or by assuming that all 

congeners were equally potent to 2,3,7,8-TCDD (USEPA, 1987a, 1989b). It appears that PCDD 

and PCDF congeners (and co-planar PCBs) act through a common mechanism of action and 

induce similar biochemical and toxicological effects. However, the toxicity of individual 

congeners was shown to vary in different bioassays, leading to a recognition of the uncertainty 

associated with the earlier approach. 

Eadon et al. (1986) described the first TEF-like method as a means to estimate potential 

human health risks associated with a PCB transformer fire in Binghamton, New York. 

Subsequently, EPA concluded that a TEF approach was the best available interim science policy 

for dealing with complex emissions of dioxins and furans from waste incineration. In 1987, EPA 

adopted an interim procedure, based on TEFs, for estimating the hazard and dose-response of 

complex mixtures containing PCDDs and PCDFs in addition to 2,3,7,8-TCDD (USEPA, 1987a). 

Following adoption of the TEF methodology in the United States and Canada, the North 

Atlantic Treaty Organization Committee on the Challenges of Modern Society (NATO/CCMS) 

examined the methodology and concluded that it was the best available interim method for 

PCDD/PCDF human health risk assessment (NATO, 1988a, b). NATO/CCMS refined the TEFs 

by including more recent data sets and more in vivo data. The NATO/CCMS panel assigned 

TEFs to OCDD and OCDF, and removed TEFs for all congeners lacking chlorine in the 2,3,7,8-

positions. Although it is theoretically possible to detect nearly all of the 210 PCDD/PCDF 

isomers in the environment, only the seventeen 2,3,7,8-substituted congeners are known to 

bioaccumulate. EPA officially adopted the revised TEFs in 1989 (TEFs-NATO89), with the 

caveat that the methodology remain interim and continued revisions be made (USEPA, 1989b; 

Kutz et al., 1990). The use of the TEF methodology for human health risk assessment and risk 
Page 86 Examination of EPA Risk Assessment Principles and Practices 

management purposes has since been formally adopted by a number of other countries (e.g., 

Canada, Germany, Italy, the Netherlands, Sweden, and the United Kingdom) (Yrjänheiki, 1992). 

The dioxin-like activity of co-planar PCBs has also been established (Safe, 1990, 1994). 

Studies show that only PCB congeners substituted in the meta and para positions are approximate 

stereoisomers of 2,3,7,8-TCDD and induce dioxin-like biochemical and toxicological effects 

(Leece et al., 1985). In 1991, EPA convened a workshop which concluded that a small subset of 

the PCBs displayed dioxin-like activity and met the criteria for TEFs (Barnes et al., 1991; 

USEPA, 1991b). 

In the years since initial adoption of the TEF methodology, additional data have been 

developed on the toxicological potency of individual PCDDs, PCDFs, and PCBs relative to 

2,3,7,8-TCDD. A joint project conducted by the World Health Organization European Centre for 

Environmental Health (WHO-ECEH) and the International Programme on Chemical Safety 

(IPCS) resulted in development of a database of all available relevant toxicological data for 

dioxin-like compounds available through 1993. Following a review of almost 1,200 peer 

reviewed publications, 146 were selected and used to calculate TEFs for PCBs (TEFs-WHO94). 

A panel of experts from eight different countries recommended interim TEFs for 13 dioxin-like 

PCBs, based on the reported results for 14 different biological and toxicological parameters, from 

a total of 60 articles (Ahlborg et al., 1994). EPA reaffirmed application of this methodology in 

human health risk assessment in its dioxin reassessment (USEPA, 2000i). 

In 1997, a second WHO-ECEH group expanded the TEF methodology to include classspecific TEFs for mammals, birds, and fish. The resulting report (van den Berg et al., 1998) 

included TEFs for 7 PCDD, 10 PCDF, and 12 PCB congeners for mammals, birds, and fish 

(TEFs-WHO98). 

In summary, the TEF concept and associated TEF factors for PCDDs, PCDFs, and PCBs 

were developed through an international consensus process that has 

a) Included input from numerous international experts, agencies, and stakeholders. 

b) Re-examined the assumptions and limitations of the concept and the TEFs. 

c) Revised TEFs as new scientific information and data became available. 

This method of adding weighted doses for related compounds is considered by the international 

scientific community to be an improvement over other options, such as (1) ignoring risks for 

non-TCDD congeners, (2) treating each congener as a separate chemical evaluation (i.e., ignoring 

the risk posed by the mixture), or (3) assigning the full toxicity of TCDD to all similar congeners. 
Examination of EPA Risk Assessment Principles and Practices Page 87 

4.5.3 Is the Additivity Assumption for TEF Justified? 

A substantial effort has been made to test the assumptions of additivity and the ability of 

the TEF approach to predict the effects of mixtures of dioxin-like chemicals. Current 

experimental evidence shows that for PCDDs, PCDFs, co-planar dioxin-like PCBs, and strictly 

AhR-mediated events, the concept of TEF additivity adequately estimates the dioxin-like toxicity 

of either synthetic mixtures or environmental extracts, despite the variations in relative 

contributions of each congener (see chapter 9 of the Dioxin Reassessment, USEPA, 2000a — 

particularly van Birgelen et al. [1994a, b] and van der Plas et al. [1999, 2000]; also Hamm et al., 

2003). 

Studies in fish and wildlife species of mixtures of PCDDs, PCDFs, and PCBs (including 

binary mixtures, synthetic complex mixtures, and environmental mixtures) support the additivity 

assumption (Zabel et al., 1995; Walker et al., 1996; Tillitt and Wright, 1997). Further, numerous 

studies that have examined the effects of environmental mixtures in marine mammals and avian 

species show a correlation between toxic effects and dietary concentrations (toxic equivalency 

quotient; TEQ) (Ross et al., 1996; Summer et al., 1996a, b; Giesy and Kannan, 1998; Restum et 

al., 1998; Shipp et al., 1998a, b; Ross, 2000). 

In summary, current experimental evidence shows that for PCDDs, PCDFs, and co-planar 

PCBs, additivity adequately estimates the dioxin-like toxicity of both synthetic mixtures and 

mixtures found in the environment. Interactions other than additivity have been observed with a 

variety of effects in both binary combinations and complex synthetic mixtures of dioxin-like and 

non-dioxin-like chemicals (commercial PCBs, PCB153). However, it appears that at these highdose exposures, multiple mechanisms of action not under the direct control of the AhR are 

responsible for these non-additive effects. 

4.5.4 What Is EPA’s Experience With TEFs? 

EPA and other organizations have estimated the risks of exposure to PCDD and PCDF 

mixtures using TEFs for more than a decade. The TEF concept and supporting literature are 

covered in detail in chapter 9 of the draft dioxin reassessment (“Toxic Equivalency Factors 

[TEF] for Dioxin and Related Compounds”) and also in the Part III, Risk Characterization. Both 

of these documents have undergone interagency review and other external peer review. The RPF 

method was also extensively reviewed outside the Agency. The Regions used the “RPF-like” 

approach developed for PAHs (USEPA, 1993a) to estimate risks posed by this group of 

compounds. The Agency used the RPF method in its assessment of the cumulative risks 

associated with exposures to organophosphorus pesticides (USEPA, 2002f, g); this effort 

underwent extensive internal and external peer review. 
Page 88 Examination of EPA Risk Assessment Principles and Practices 

 In 1998, EPA and the Department of the Interior sponsored a workshop that examined 

application of the TEF concept and supporting assumptions using ecological risk assessment case 

studies. The resulting document, Workshop Report on the Application of 2,3,7,8-TCDD Toxicity 

Equivalence Factors to Fish and Wildlife (USEPA, 2001e), provides general conclusions 

regarding application of and uncertainties associated with the TEF methodology in ecological 

risk assessment, as follows: 

a)	 The TEF methodology is technically appropriate for evaluating risks to fish, birds, 

and mammals associated with AhR agonists, and it can support risk analyses 

beyond screening-level assessments. 

b)	 The methodology entails less uncertainty and is less likely to underestimate risks 

than are methods based on single compounds (i.e., 2,3,7,8-TCDD). 

c)	 Because total PCBs in the environment can consist of many compounds that vary 

in concentration and potency as AhR agonists, the TEF methodology provides a 

means for accounting for their variable potency. 

d)	 The uncertainties associated with using the methodology are not thought to be 

larger than other sources of uncertainty within the ecological risk assessment 

process. 

For human health assessment, issues concerning TEFs are being addressed through the 

dioxin reassessment, its review by the NAS, and the policy implementation document planned 

for after the release of the reassessment. In the interim, the Agency is obliged in many instances 

to use the best available science and risk assessment methods in regulatory decisions: hence the 

movement to use the internationally accepted practice of TEQ calculations for PCBs. From an 

international scientific perspective, the current TEQ concept and values are a necessary and 

appropriate interim procedure in advance of the development of more refined methods (van den 

Berg et al., 1998; USEPA, 2001e). 

4.6	 Model Use 

4.6.1	 Why Does EPA Use Environmental Models? 

A wide variety of models have been used over the years to support environmental 

decision making by the Agency. These models often involve environmental characterization, 

environmental fate and transport, contact between stressors and receptor populations, 

relationships between exposure and adverse human health and ecological effects, quantitative 

estimation of risk, and economic impact models, all of which affect the manner in which the 

Agency chooses to address a multitude of environmental questions. Frequently, modeled 
Examination of EPA Risk Assessment Principles and Practices Page 89 

estimates become the basis for environmental cleanup, protection, or regulation. Environmental 

models are vital to the Agency’s mission because they allow us to (1) predict the effects of future 

actions, (2) estimate exposures and risks where no measured data are available, and (3) attribute 

risks to individual sources. Models, therefore, underlie how the Agency chooses to address a 

multitude of environmental questions. At issue is whether the underlying assumptions, 

algorithms, etc., used in the modeling — and potentially the analytic structure of the models 

themselves — are articulated to the public well enough for stakeholders to assess the validity of 

modeling assessments. Further, analysis of the uncertainties in modeled estimates should be 

clear and transparent to those unfamiliar with the models. 

4.6.2 How Does EPA Approach Environmental Models? 

As the Agency’s scientific understanding of environmental issues has improved over the 

years, the models we use to understand them have become increasingly complex, requiring the 

involvement of a wide range of scientific disciplines in the development of the models as well as 

in the interpretation of the results. As a consequence, the Agency and others have become 

increasingly focused on consistency, quality, and duplication of effort in model development, 

selection, and application. Such issues can, unfortunately, lead to risk characterizations that are 

not “transparent.” Transparency requires explicitness in the presentation of each step in the risk 

assessment process. It ensures that any reader understands all the steps, logic, key assumptions, 

limitations, and decisions in the risk assessment, and comprehends the supporting rationale that 

lead to the outcome, including any modeling that is done for fate/transport and exposures. 

Generally, EPA risk assessments address several categories of uncertainty. Each of these 

merits consideration. For example, for models, there are uncertainties associated with the 

selection of specific scientific models for each parameter in risk assessment, e.g., dose-response 

models, models of environmental fate and transport, and exposure models. The selection of 

parameters for these models creates even more sources of uncertainty. Ultimately, EPA guidance 

recommends that risk assessors identify those uncertainties that, if changed, would substantially 

affect the modeled outcomes. 

4.6.3 What Is EPA Doing To Improve the Use of Environmental Models? 

As early as 1989, EPA’s SAB expressed its concerns on EPA’s use of environmental 

models in Resolution on Use of Mathematical Models by EPA for Regulatory Assessment and 

Decision Making (USEPA, 1989c). Since that time, the SAB has continued to recommend that 

EPA “establish a general model evaluation protocol, provide sufficient resources to test and 

confirm models with appropriate field and laboratory data, and establish an Agency-wide task 

group to assess and guide model use by EPA.” 
Page 90 Examination of EPA Risk Assessment Principles and Practices 

In response to these concerns, the Deputy Administrator established the Agency Task 

Force on Environmental Regulatory Modeling (ATFERM) in 1992. In 1994, the ATFERM 

published the Report of the Agency Task Force on Environmental Regulatory Modeling — 

Guidance, Support Needs, Draft Criteria and Charter (USEPA, 1994b). The report concludes 

that (1) there is a need for training and technical support, (2) there is a need for model use 

acceptability criteria, (3) there is a need for Agency guidance for conducting external peer review 

of environmental regulatory modeling, and (4) there is a need for a Committee on Regulatory 

Environmental Modeling (CREM). To this end, the report included a section entitled “Guidance 

for Conducting External Peer Review of Environmental Regulatory Modeling.” 

In 1994, the EPA Risk Assessment Forum developed a document entitled Model 

Validation for Predictive Exposure Assessments (USEPA, 1994c), a draft protocol for model 

validation that defined a set of procedures for evaluating models for exposure assessments. In 

1997, EPA’s Officer of Research and Development (ORD) and program offices conducted an 

Agency-wide conference (called the Models-2000 Workshop) to facilitate Agency adherence to 

existing guidance on modeling, to define and implement improvements to the way in which the 

Agency develops and uses models, and to recommend an implementation and improvement plan 

for enhancing modeling within the Agency. As a follow-on to the workshop, 10 Action Teams 

and a Steering/Implementation Team were established to develop further action plans for 

improving the Agency’s use of models. 

In 1999, a White Paper on the Nature and Scope of Issues on Adoption of Model Use 

Acceptability Guidance was developed (USEPA, 1999c). The paper, by discussing current 

practices and case studies, reviewed progress in and identified issues associated with developing 

model use acceptability guidance. 

In February 2000, the Deputy Administrator established the CREM and developed the 

framework of guiding principles for its activities. The CREM was established to promote 

consistency and consensus within the Agency on mathematical modeling issues (including 

modeling guidance, development, and application) and to enhance both internal and external 

communications on modeling activities. The CREM supports and enhances existing modeling 

activities by Agency program offices. The CREM provides the Agency with consistent yet 

flexible modeling tools to support environmental decision making, in particular as they relate to 

the development and implementation of programs with cross-Agency implications. Further, the 

CREM provides EPA staff and the public with a central point for inquiring about EPA’s use of 

modeling. 

In May 2000, the CREM initiated several cross-Agency activities that were designed to 

enhance the Agency’s development, use, and selection of regulatory environmental models. One 

of these activities involved the ultimate development of modeling guidance. The CREM 

determined that a workshop to facilitate discussion of good modeling practices among 
Examination of EPA Risk Assessment Principles and Practices Page 91 

participants would constitute a viable starting point for developing modeling guidance. To 

organize the workshop and develop issue papers for discussion, a cross-Agency Model 

Evaluation Action Team was formed in June 2000. Over the next several months, the Action 

Team (1) reviewed the White Paper on the Nature and Scope of Issues on Adoption of Model Use 

Acceptability Guidance to identify model evaluation issues; (2) developed a list of components, 

or elements, to follow in evaluating a regulatory environmental model; (3) identified several 

general principles to support the use of model evaluation guidance; (4) developed an 

organizational plan for the workshop that covered the identified model evaluation elements; (5) 

held the workshop; and (6) developed a summary report of the workshop. 

In February 2003, EPA’s Administrator directed the CREM to draft Agency guidance on 

environmental models and create a Web-accessible inventory of EPA’s most frequently used 

models. Both these efforts address the transparency and uncertainty with which environmental 

models — including fate and transport models — are used to make decisions. First, in its 

guidance on environmental models, EPA intends to recommend that the modeling community 

document all steps taken to develop and evaluate models; the guidance will also recommend best 

practices for evaluating models. Second, EPA intends to house this documentation of model 

development and evaluation within the Web-accessible inventory of models. 

In addition to the activities noted above, EPA issued Order 5360.1 A2 (USEPA, 2000c), 

which covers collection and use of environmental data, including information produced from 

models in the areas of (1) use of a systematic planning approach to develop acceptance or 

performance criteria; (2) approved Quality Assurance Project Plans (QAPPs) or equivalent 

documents; and (3) assessment of existing data used to support Agency decisions or other 

secondary purposes (to verify that they are of sufficient quantity and adequate quality for their 

intended use). Requirements for QAPPs are provided in EPA QA Manual 5360 A1 (USEPA, 

2000d) for EPA personnel and EPA QA/R-5 (USEPA, 2001f) for extramural personnel. 

The atmospheric dispersion models we use (e.g., the Industrial Source Complex model, or 

ISC, and AERMOD) have been through rigorous scientific peer review. To improve our tools 

for modeling fate and transport from air into other media, EPA developed the Fate, Transport and 

Ecological Exposure Module of the Total Risk Integrated Methodology, or the TRIM.FaTE 

model. TRIM.FaTE not only handles the partitioning of pollutants between environmental 

media, but does so in an interactive way, allowing pollutants to move back and forth between 

media throughout the simulation according to the simulation conditions. The input values, any 

assumptions, and transfer algorithms are completely transparent and accessible to the user 

because they are stored in an electronic “library” accessed by the model, not “hardwired” into the 

model itself. This feature allows the Agency to easily replace algorithms or values as new 

findings indicate they are appropriate, or as appropriate for different applications. A variety of 

documents have been prepared on the model and algorithms compiled in the library for 

applications to date (e.g., USEPA, 2002h, i, j); these documents are available along with the 
Page 92 Examination of EPA Risk Assessment Principles and Practices 

model on the Agency’s Web site. TRIM.FaTE also includes features for sensitivity and Monte 

Carlo analyses to assist in developing analyses of uncertainty and variability for multimedia risk 

assessments. 

With regard to inhalation exposure models, EPA has developed methods that use human 

time activity patterns and stochastic sampling to derive probabilistic activity-based estimates of 

exposure. The current example of this is the Human Exposure Event module of TRIM (the 

TRIM.Expo inhalation model, also known as the air pollutant exposure model or APEX). This 

model is derived from the probabilistic National Ambient Air Quality Standards exposure model 

(pNEM) and has been developed for both criteria and HAP applications. TRIM.Expo is 

available on the EPA Web site and by its design it relies on data from input files that are 

completely transparent to the user. 

Regarding uncertainty analysis, each Agency risk assessment needs to include an analysis 

of important uncertainties that, if reduced, might change the outcome of the assessment (see the 

Agency’s 1997 Policy for Use of Probabilistic Analysis in Risk Assessment and Guiding 

Principles for Monte Carlo Analysis). The Agency believes that it is generally preferred that 

quantitative uncertainty analyses be presented in each risk characterization, but there is no single 

recognized guidance on how to conduct such uncertainty analyses. Even if the uncertainty 

analysis is conducted qualitatively, the Agency still considers it to be of great value to a risk 

manager. 

4.6.4	 How Does EPA Approach Fate and Transport Modeling and Account for 

Uncertainty? 

Once chemicals are released to the environment, they are subject to a wide range of 

environmental processes that influence their fate and transport. These processes may affect their 

location (e.g., meteorology, wind, water currents), their chemical structure (e.g., transformation, 

reactivity), and the media in which they partition or reside. These processes’ level of influence 

varies significantly, often depending on the chemicals’ physical and chemical properties as well 

as the conditions they encounter in the environment. 

This section describes the modeling conducted under OPP to illustrate EPA’s approach to 

fate and transport modeling. (Fate and transport modeling in OPP efforts typically encounters 

multiple media and therefore is a good example of broad modeling practices used by EPA.) Fate 

and transport models are used in OPP to estimate ambient concentrations for ecological risk 

assessment and drinking water concentrations for human risk assessments. For ecological risk 

assessments, estimated environmental concentrations are divided by a toxicity reference value to 

yield an index that is compared to a level established to constitute an acceptable risk. For human 

risk assessments, estimated drinking water concentrations are incorporated into a total dietary 

assessment, which is then evaluated for risk. 
Examination of EPA Risk Assessment Principles and Practices Page 93 

EPA responds to uncertainty in these assessments depending on the type of decision to be 

made. In estimating pesticide concentrations in water resources, OPP uses a tiered screening 

approach (consistent with the general approach discussed in chapter 2) that involves a 

combination of modeling and analysis of monitoring data. In OPP’s tiered modeling system, the 

program begins with a public-health-protective, national, highly vulnerable estimate and then, if 

concerns exist, progresses to regional and eventually local, probabilistic, more resource-intensive 

estimates. This triage approach is designed to provide a thorough analysis of each pesticide, 

while focusing OPP’s efforts on those pesticides that pose the greatest potential risk. 

The intent of the screening approach is to estimate pesticide concentrations in water from 

sites that are vulnerable to runoff or leaching so that the program can be confident that any 

pesticide that passes the screening tiers poses a low possibility of harming human health, 

wildlife, or the environment. Therefore, during these screening assessments, the intent is to 

address the uncertainty regarding the most sensitive scenario that may exist in areas of pesticide 

application. 

These initial screening assessments establish the environmental setting and input 

parameters that would result in a high-end exposure of a pesticide after application. Half-life 

data from laboratory studies serve as inputs, as described in the publicly available input 

parameter guidance. The initial screening levels models used by OPP are: for the surface water 

pathway, GENEEC2 (GENeric Estimated Exposure Concentration) and FIRST (FQPA Index 

Reservoir Screening Tool), and for the groundwater pathway, SCI-GROW (Screening 

Concentration in GROund Water). 

If a pesticide raises a concern after the initial screening tier, a set of higher-tiered surface 

water models called PRZM/EXAMS are used to conduct a more refined assessment. These 

models, developed by ORD, allow the Agency to focus on specific sites (by accounting for sitespecific properties including soil type, weather, and agricultural practices) or watersheds in its 

analysis. At the present time, OPP does not have a higher-tiered groundwater model. 

A tier II or PRZM/EXAMS refined water assessment uses a region-specific site with an 

agriculturally relevant soil that is vulnerable to runoff and is of a large extent in the watershed 

(i.e., a benchmark soil). The assessment also uses 30 years of monitored weather data for the site 

and maximum pesticide use rates. The combination of site characteristics and fate property 

inputs is designed to approximate a “high-end” exposure scenario. A tier II drinking water 

assessments uses an index reservoir drinking water source (USEPA, 1998d), and ecological 

exposures assessments are conducted using a small fixed-volume water body, the “farm pond.” 

The drinking water reservoir modeled is Shipman City Lake in Illinois; the farm pond is located 

in Georgia. Drinking water estimates are adjusted to reflect the amount that a major crop, 

specifically corn, wheat, cotton, or soybeans (USEPA, 1999d), may be grown in a watershed 

using a national percent crop area factor. Crop area factors for other crops or combinations of 
Page 94 Examination of EPA Risk Assessment Principles and Practices 

crops to which a pesticide may be applied in the same watershed are assigned a national default 

value representing the highest percent of a watershed in agricultural production. 

Recently, region-specific default percent crop areas were developed to reflect different 

cropping practices across the United States. Estimates may be further refined by focusing on 

additional region-specific use sites and typical usage and application practices. Estimated 

concentrations from PRZM/EXAMS in the initial tiers are a 1 in 10 year occurrence frequency 

and, for cancer risk assessments, a 30-year overall mean. Additional refinements are pesticidespecific and may be performed at the request of risk managers seeking to focus on risk 

mitigation. These refinements may include a probabilistic dietary assessment using the total 

PRZM/EXAMS output of 30 years of daily values. 

If degradates are observed in the laboratory studies that are of toxicological significance, 

have or are predicted to have pesticidal activity based on structure, or are a significant component 

of the residues in an environmental compartment, modeling is conducted on these degradates. 

The screening-level models are bypassed in favor of PRZM EXAMS to model degradates and are 

assessed in separate modeling runs. The amount of data available to serve as inputs varies. Halflives can sometimes be obtained from the registrant-submitted studies if those studies were 

conducted for a period greater than the half-life of the parent. Defaults are used in the absence of 

data, as described in the input parameter guidance. 

4.6.5	 Does EPA Consider or Review Existing Data on Environmental 

Concentrations? 

Monitoring data are preferred when estimating concentrations; however, monitoring data 

are generally either not available or inadequate. Monitoring data are not available for new, 

never-registered chemicals. No or little monitoring data will be available for a chemical that has 

little use in the United States or has a “niche” market. In some cases, after reviewing the 

monitoring data, modeling outputs are selected because monitoring was conducted in areas in 

which the chemical was not used or had little use. Generally, monitoring data in targeted 

pesticide use areas are most useful for estimating long-term time-weighted concentrations and 

potential chronic impacts. Low sample frequency and lack of statistically representative use sites 

limits the usefulness of monitoring data for estimating peak concentrations and its potential acute 

impacts. Sufficient data are sometimes available for widely used pesticides, pesticides that are in 

multi-residue monitoring methods, or pesticides that have the interest of the public. In all risk 

assessments, available monitoring results are compared to screening-level modeling estimates. 
Examination of EPA Risk Assessment Principles and Practices Page 95 

4.6.6	 Does EPA Consider and Use Case-Specific Data in Models? (Or: How Does 

EPA Decide To Modify Default Assumptions in Later Tiers Using Models?) 

As in all aspects of risk assessment, EPA prefers to use specific data rather than default 

assumptions when working with models. Default assumptions are used when there are data gaps. 

When data of sufficient quality are available, EPA uses such data. For example, the pesticide 

industry is responsible for submitting the data that serve as the inputs into these models. Data 

requirements are described in 40 CFR Part 158, and the protocols for conducting these studies 

are discussed in the Pesticide Assessment Guidelines. Registrants may use Organisation for 

Economic Cooperation and Development (OECD) guidelines or other guidance provided the data 

meets the purpose of the study. In selecting model parameters, OPP established publicly 

available guidance for selecting inputs (Water Models, Guidance for Selecting Input Parameters). 

The amount of data available determines how confident an input will become. 

4.6.7	 How Will EPA Improve Pesticide Fate and Transport Modeling in the 

Future? 

The Agency’s goal is to develop realistic estimates of pesticide concentrations in ambient 

and drinking water. The following efforts are underway to improve its tools and technology: 

a)	 Ground water: A tier II model that generates daily or short-time-step 

concentrations for use in a probabilistic dietary assessment will be in use in late 

2004. Currently available groundwater models will be assessed jointly with the 

Pest Management Regulatory Agency/Health Canada. In addition, a refined 

version of SCI/GROW is being investigated that uses all of the data from 

prospective groundwater studies submitted by the pesticide industry. 

b)	 Surface water: Modifications are being made to PRZM/EXAMS to allow the size 

of the receiving body to be adjusted to reflect the size of a typical pond or 

reservoir in the area being modeled. A geographic information system is being 

developed that maps all of the drinking water intakes in the United States and 

their associated watersheds. Additional data layers will be added to allow the 

Agency to model specific sites and their cropping patterns. Spray drift is a fixed 

variable in PRZM/EXAMS. Spray drift models will be linked to PRZM/EXAMS 

to generate realistic drift inputs (USEPA, 1997e, 1999e). Watershed-scale 

regression models are being developed in cooperation with the U.S. Geological 

Survey to estimate a distribution of concentrations according to watershed 

properties and pesticide usage areas (Larson et al., in press; Larson and Gilliom, 

2001). 
Page 96 Examination of EPA Risk Assessment Principles and Practices 

c)	 Drinking water treatment: A workshop planned for January 2004 will address the 

types of data needed to be able to factor drinking water treatment into estimates of 

pesticide concentrations. Additional workshops and a Federal Insecticide, 

Fungicide, and Rodenticide Act (FIFRA) Scientific Advisory Panel meeting are 

expected to finalize any new data requirements that may be imposed on industry 

to provide data. 

d)	 Tiered water exposure assessment process: Interaction is conducted with the 

human health risk assessors and risk managers in OPP to standardize the 

modeling tiers and generate the data needed for conducting probabilistic exposure 

assessments earlier in the process and to identify the tools and technologies to 

routinely conduct probabilistic analyses. The program’s goal is to account for 

daily and seasonal variations in residues over time due to the time of application 

and runoff/leaching events, year-to-year variations in weather patterns, and placeto-place variability in residues caused by the water source and regional/local 

factors (such as soil, geology, hydrology, climate, crops, pest pressures, and 

usage) that affect the vulnerability of a source. 

4.6.8	 Does EPA Use Worst-Case Assumptions in Modeling That Are Unreasonable 

and Do Not Account for Degradation, Partitioning, and “Sinks” in the 

Environment? 

Default assumptions are frequently used in modeling, specifically in exposure assessment 

for each medium or cumulative assessment. Model inputs for fate and transport use the most 

sensitive soil type, combined with the lowest rates of degradation, air exchange, or water 

dispersion, and 100% of mass used in each scenario. These default assumptions are used to 

address data gaps and to ensure that risks will not be underestimated given those uncertainties. 

The degree to which defaults are incorporated into any modeling exercise depends on the 

purpose of the assessment and the availability of data. In a screening assessment with little or no 

available data, several default assumptions may be used, such as the assumption that the 

chemical is stable and not subject to transformation or degradation. In fact, the potential for 

degradation of pesticides in the environment is and has been incorporated into many aspects of 

OPP risk assessments, as illustrated in the examples below. 

a)	 Food: In dietary exposure and risk assessments, the use of monitoring data (e.g., 

the U.S. Department of Agriculture’s Pesticide Data Program or market basket 

surveys) incorporates potential degradation of pesticides from “farm gate” 

residues to “dinner plate” residues. Dietary exposure estimates include 

consideration of the impact of typical consumer practices such as washing and 

peeling, and commercial practices during processing (juicing, blanching, freezing, 
Examination of EPA Risk Assessment Principles and Practices Page 97 

etc.), that may reduce or concentrate residues. For pesticides found largely on the 

surfaces of fruits and vegetables, considerable reduction of residues has been 

observed in processing studies, as well as in monitoring data when results for 

fresh and processed commodities can be compared. Concentration of residues is 

observed more often in dried commodities. 

b)	 Water: In determining the potential concern for pesticide residues in drinking 

water, OPP risk assessments incorporate modeled surface and groundwater values 

generated using chemical-specific environmental fate inputs (e.g., hydrolysis, 

anaerobic and aerobic metabolism, and microbial degradation). If sufficient data 

are available, assessments also consider the impact of treatment on potential 

residues in drinking water. Finally, monitoring data are evaluated for use in the 

risk assessment. (If used quantitatively, then degradation in the environment is 

inherently considered; if used qualitatively, a comparison between modeled and 

monitored values is provided as characterization.) Currently, the impact of water 

treatment on drinking water concentrations is being assessed qualitatively. 

Following a recommendation from the FIFRA Scientific Advisory Panel meeting 

during September 2000 (USEPA, 2000j), ambient concentrations are used in a 

drinking water assessment unless there are data on the impact of water treatment. 

c)	 Residential: Exposure and risk estimates for pesticides in residential settings are 

generated using both default assumptions and chemical-specific data. The default 

assumptions are used largely in screening-level assessments. Chemical-specific or 

activity-based information is incorporated as available. The timing and duration 

of exposure are the most critical factors in estimating risk. For example, in 

estimating toddlers’ dermal exposure to pets, the highest residues on the pets’ fur 

(and the highest exposure to the child) are expected to occur on the day of 

application. For longer-term exposure durations, the decline in residues on fur is 

taken into consideration. In assessing dermal exposure to golfers from treated 

turf, short-term non-cancer risks are estimated on the day of application at the 

label rate, but cancer risks are estimated based on average application rates, and 

take into account the frequency of golfing over a lifetime. Exposure assessments 

that consider “sinks” (e.g., carpeting or soil) can be problematic due to data 

deficiencies, but there is an assumption that less pesticide would be available for 

dermal exposure, and an adjustment can be made in calculating the risks using, for 

example, a “soft surface” exposure factor. In estimating indoor inhalation 

exposures, the registered label is used as the basis for determining the re-entry 

time and ambient concentrations, but breathing rate and level of activity 

considerations can be included in the final risk estimates. 
Page 98 Examination of EPA Risk Assessment Principles and Practices 

d)	 Occupational: For occupational handlers (those who mix, load, and/or apply 

pesticides), estimated exposures are driven by pesticide formulation type, the 

application method (e.g., aerial or ground boom), and the number of acres treated 

per day. Exposure and risk estimates are calculated assuming different levels of 

protective clothing and using the geometric mean of exposures (unit exposures 

based on real data conducted with a variety of formulation types and application 

equipment). Dermal risks estimated for different clothing types adjust for 

potential protection afforded by a second layer of clothing or gloves, and consider 

how much of the pesticide may actually get on the skin. Inhalation risks 

calculated using unit exposure data already take into account activity levels, 

breathing rate, and the potential ambient air concentrations for various pesticide 

formulations during application. For harvesting or other foliar-contact activities 

undertaken following application, the degradation of pesticide residues on leaves 

over time is considered and used in the calculations. For estimating cancer risks, 

exposures are calculated based on typical application rates, then amortized over a 

lifetime. 

In all of the above, the best available data are used to estimate exposure and risk and, to the 

extent possible, reflect the use pattern for specific pesticides. If risks exceed the level of concern, 

more specific information about a particular pesticide/crop/pest combination may be obtained to 

provide more refined estimates of exposure and risk. 
Examination of EPA Risk Assessment Principles and Practices Page 99 

5. SITE- AND CHEMICAL-SPECIFIC ASSESSMENTS 

5.1 Overview 

This chapter addresses EPA’s use of site- or chemical-specific information in risk 

assessments and the default assumptions that are often used in the absence of such information. 

Many comments that EPA received regarding site- and chemical-specific information pertain to 

assessments concerning waste sites. Therefore, this chapter largely focuses on the comments in 

the context of human health risk assessment guidance for Superfund, which is also followed by 

the Resource Conservation and Recovery Act Corrective Action program, as appropriate. EPA 

also received a number of comments directed at risk management decisions made at specific 

Superfund sites. Indeed, many of the site-specific comments repeat comments filed with the 

Agency either during the development of the proposed plans or during the public comment 

period for proposed plans at those specific sites. The EPA officials responsible for implementing 

the Comprehensive Environmental Response, Compensation, and Liability Act (CERCLA) at 

those sites either have already responded to those comments or will be responding to them 

through the CERCLA process. Therefore, this document does not respond to comments on 

specific sites. Further information about the role of site- and chemical-specific data in other 

regulatory programs may be available from individual program offices (refer to the EPA home 

page at http://www.epa.gov/epahome/programs.htm) and by other examples found elsewhere in 

this document (e.g., the “porch potato”). 

Due to the uniqueness of individual waste sites, site-specific information plays an 

important role in risk assessments and management decisions on a regional level. In general, 

EPA considers site- and chemical-specific information in risk assessments when it is available 

and appropriate, then uses default assumptions when data gaps exist. Default assumptions are 

typically based on scientific, peer reviewed data and aid in the evaluation of aspects such as 

various age groups and various populations (e.g., residents, workers, Native American tribes, 

subsistence fishers). 

In terms of program goals and protecting public health, it is appropriate for EPA to use 

default assumptions when data are not available to use in a risk assessment. Note that the default 

exposure values are not all high-end, are not “worst-case,” and have data supporting their 

derivation (in addition to being peer reviewed). The following table presents examples of 

commonly used default exposure factors and their associated percentile distribution, which range 

from median to 95th percentile. 
Page 100 Examination of EPA Risk Assessment Principles and Practices 

Table 5-1: Examples of Default Exposure Values With Percentiles 

Exposure Pathway Percentile Source of Data 

Drinking water 

consumption: 

2 liters/day 

90th Approximately a 90th percentile value (USEPA, 

2000k). 

Soil ingestion rate for 

children: 

200 mg/day 

65th Analyses and distributions constructed by Stanek 

and Calabrese (1995a, b, 2000) places the 200 mg 

ingestion rate around the 65th percentile of average 

daily intakes throughout the year. The Stanek and 

Calabrese analyses suggests that ingestion rates for 

children in the top 10% (i.e., the high end) of the 

distribution would be greater than 1,000 mg/day. 

Residence duration: 

30 years 

90th 

80th 

90th–95th 

For home owners, farms, and rural populations; 30 

years is greater than the 95th percentile residence 

time for renters and urban populations. 

Body weight: 

70 kg 

50th For males and females 18 to 75 years old (NCHS, 

1987) 

Risk assessments and the collection of data and other information are conducted on a sitespecific basis by the EPA regional offices, within the regulatory requirements of the waste 

programs. Public participation is a critical component of the regional Superfund process. For 

example, public knowledge about the history of disposal practices at an abandoned waste site 

often informs the conceptual site model design and development of sampling plans around the 

site. Another example of site-specific information involves people’s interaction with their 

environment. Some populations may eat more fish than would be estimated using EPA default 

exposure values based on differences in consumption patterns. If resources are available, and 

considering the population and other factors, a site-specific survey may be conducted to assess 

the consumption patterns for the specific population. 

Often, when EPA actively seeks public participation, the Agency staff also need to clarify 

what constitutes useful data for risk assessment. For example, a soil sample placed in a 

household container does not provide useful or adequate data for a risk assessment. EPA would 

explain that information about the extent of contamination is collected following the quality 

assurance/quality control guidance (USEPA, 2000l) using a standard protocol with quality 

control and quality assurance procedures. 

To address the concerns and comments about use of site- and chemical-specific data and 

to fully understand assessments of cancer risks and non-cancer health hazards, it is important to 
Examination of EPA Risk Assessment Principles and Practices Page 101 

understand their underpinnings. Among these are the regulatory requirements and purpose of the 

programs for which the assessments are conducted, whom the programs seek to protect, the role 

of default assumptions and site- and chemical-specific information in risk assessments, the types 

of risk assessment performed, and factors considered in risk management decisions. The 

following sections present information about these aspects of the Superfund program. 

5.1.1	 What Are the Superfund Program’s Purpose and Regulatory Requirements? 

CERCLA, 42 U.S.C. 9605 (CERCLA, 1980), as amended by the Superfund Amendments 

and Reauthorization Act of 1986 (SARA), P.L. 99-499 (SARA, 1986), requires that actions 

selected to remedy hazardous waste sites be protective of human health and the environment. 

The National Oil and Hazardous Substances Pollution Contingency Plan, or NCP, is the 

regulation that implements CERCLA (NCP, 1990a, b). The NCP establishes the overall 

approach for determining appropriate remedial action at Superfund sites (NCP, 1990a, b). 

The national goal of the Superfund program is to select remedies “ ... that are protective 

of human health and the environment, that maintains protection over time, and that minimizes 

untreated waste”(NCP, 1990a). This means that the Superfund program needs to protect human 

health and the environment from current and potential future threats posed by uncontrolled 

hazardous substance releases. Pursuant to the NCP, decisions at Superfund sites involve 

consideration of site-specific information and cancer risks and non-cancer health hazards 

associated with both current and future land use conditions. 

5.1.2	 Whom Does the Superfund Program Seek To Protect? The Reasonable 

Maximum Exposure Scenario 

One of the policy goals of the Superfund program is to protect a high-end, but not worstcase, individual exposure: the reasonable maximum exposure (RME). The RME is the highest 

exposure that is reasonably expected to occur at a Superfund site. As described in the preamble 

to the NCP (NCP, 1990a, b), the RME will 

... result in an overall exposure estimate that is conservative but within a realistic range of exposure. Under 

this policy, EPA defines “reasonable maximum” such that only potential exposures that are likely to occur 

will be included in the assessment of exposures. The Superfund program has always designed its remedies 

to be pro tective o f all individ uals and enviro nme ntal rec epto rs that ma y be ex pos ed at a site; con sequ ently, 

EP A be lieves it is imp ortan t to includ e all reas ona bly exp ected expo sures in its risk assessm ents ... 

In addition to evaluating the risks to the RME individual, EPA evaluates risks for the 

central tendency exposure (CTE) estimate, or average exposed individual. This approach is 

consistent with the Risk Characterization Policy and Handbook (USEPA, 1995a, 2000a). CTE 

estimates give the risk manager additional information to consider while making decisions at a 
Page 102 Examination of EPA Risk Assessment Principles and Practices 

site. Pursuant to the NCP, decisions at Superfund sites are based on cancer risks and non-cancer 

health hazards associated with RME estimates under both current and future land use conditions. 

5.1.3 Is the RME Overly Conservative? 

EPA received comments implying that the RME represents a “worst-case” or “overly 

conservative” exposure estimate. As described above, the RME is not a worst-case estimate — 

the latter would be based on an assumption that the person is exposed for his or her entire 

lifetime at the site. As the preamble to the NCP (NCP, 1990a, b) states: 

... T he reasonable maximum exposure scenario is “reasonable” because it is a product of factors, such as 

concentration and exposure frequency and duration, that are an appropriate mix of values that reflect 

averages and 95th percentile distribution ... 

The RME represents an exposure scenario within the realistic range of exposure, since the 

goal of the Superfund program is to protect against high-end, not average, exposures. The “high 

end” is defined as that part of the exposure distribution that is above the 90th percentile, but 

below the 99.9th percentile. The approach was developed with technical support from ORD and 

is consistent with the peer reviewed EPA Exposure Assessment Guidelines (USEPA, 1992a). 

Further, for some individual activity patterns, the RME may be modified based on 

site-specific considerations (e.g., workers who may remain in one location, individuals who live 

in a residence for their lifetimes, Native Americans remaining on tribal lands, children exhibiting 

pica behavior). The Superfund program supports the development of reasonable risk assessments 

that address the exposure and risk to all segments of the community, not only the “average” 

individual. EPA seeks to protect “sensitive populations,” segments of the general population that 

are at greater risk, either because of particular sensitivity to the toxic effects of certain chemicals 

or because they experience higher exposures than the general population, as children do. Under 

the NCP (NCP, 1990a, b), the Superfund program achieves this goal of protecting public health 

by using the RME approach. 

5.2 Superfund Risk Assessment Guidance 

5.2.1 What Are Some Types of Risk Assessment Under Superfund? 

EPA conducts risk assessments that vary in length and degree of detail, from screeninglevel assessments to comprehensive site-specific baseline assessments of hazardous waste sites. 

Risk assessments may focus on chronic, long-term exposures and/or evaluation of acute 

exposures that may require an emergency response. They are often conducted using a triage 

approach, beginning with a screening-level assessment to determine if a more comprehensive 

assessment is necessary (USEPA, 1988b, 1996a, 1999a, 2001b, 2002k). 
Examination of EPA Risk Assessment Principles and Practices Page 103 

5.2.2	 How Does Superfund Use Site-Specific Concentration and Maximum 

Concentration Data in Risk Assessment? 

When evaluating the information used in risk assessment, it is important to consider the 

level of risk assessment being conducted. In screening-level assessments, risk assessors may 

assume that receptors are in contact with the maximum concentration for the entire exposure 

duration to place an upper bound on exposure estimates. This approach is used to ascertain 

which exposure pathways or contaminants are most important, so that resources may be more 

efficiently used in developing a comprehensive risk assessment for the most important 

contaminants and exposure pathways. 

The concentration term is a sensitive parameter in an exposure assessment. In most risk 

assessments, an estimate of the average chemical concentration is used to represent the exposure 

concentration or exposure point concentration (EPC) for assessing long-term or chronic exposure 

scenarios. Because of the uncertainty associated with estimating the true average concentration, 

the 95% upper confidence limit of the arithmetic mean (UCL95) is used to account for 

uncertainties (USEPA, 1992b, 2002l). 

The UCL95 is highly dependent on the number of samples collected. The greater the 

number of environmental samples, the closer the UCL95 will be to the true mean. Sometimes, 

environmental data are limited or there is extreme variability in measured or modeled data. 

When this is true, the upper confidence limit can be greater than the highest measured or 

modeled concentration and neither the average nor UCL95 is a reliable estimate of the exposure 

point concentration. If additional data cannot practicably be obtained, it is more appropriate to 

use the highest measured or modeled value as the EPC. In cases where the true mean is actually 

higher than this maximum value, especially if sampling data are very limited, the maximum 

concentration is used in the calculation (USEPA, 1992b). 

Site-specific contaminant data are collected for all sites. In some cases, the maximum 

concentration is used when data are limited or the calculated UCL95 exceeds the maximum 

concentration. However, the maximum concentration is rarely used for hazardous waste sites in 

which a comprehensive baseline risk assessment is being conducted, since reliable site-specific 

data are typically collected as part of the remedial investigation (RI) process. During the 

planning and scoping phase, the minimum number of samples that is sufficient to characterize 

environmental media is determined using EPA’s data quality objectives process (USEPA, 2000l). 

The reality is that the number of samples collected from each medium, and the resulting certainty 

achieved, need to be balanced by the resources available and the purpose of the assessment. 

Lastly, it is important to note that it is appropriate to assume individuals are exposed to 

the maximum concentration when acute exposure scenarios are being evaluated, such as those 

conducted for emergency removal actions. Another example: when “hot spots” with relatively 
Page 104 Examination of EPA Risk Assessment Principles and Practices 

high concentrations are present in surface soil, then acute toxicity is of most concern and the 

focus is on measuring or estimating short-term, peak environmental concentrations. 

5.2.3	 Has the Superfund Guidance Been Updated and Externally Peer Reviewed 

Since It Was First Released in 1989? 

EPA continues to develop and peer review guidance to assist risk assessors conducting 

various types of risk assessments in Superfund. Since the development of the Superfund Public 

Health Evaluation Manual (USEPA, 1986c), and later in 1989 the RAGS Part A, Human Health 

Evaluation Manual (USEPA, 1989a), the Superfund program developed appropriate guidance to 

help risk assessors evaluate risks to the RME individual (USEPA, 1996a, 1999a, f, 2001a, b, g, 

2002k). The guidance consistently included internal review by risk assessors across the Agency, 

external comment by stakeholders, and external peer review where appropriate based on our peer 

review guidance. 

External peer reviews have been conducted for the majority of guidance documents 

developed since 1989. For example, as early as 1989, the RAGS Part A was reviewed by the 

Agency’s Science Advisory Board. In addition, since the issuance of the Agency policy on peer 

review in 1994 (USEPA, 1994d) and subsequent handbook (USEPA, 2000b), where appropriate 

and consistent with peer review policies, Superfund Risk Assessment guidance has undergone 

external peer review. The external peer review meetings have included opportunities for the 

stakeholders to participate through public comment. All guidance documents are made available 

on the Internet for public comment and addressed in the final documents. 

The probabilistic risk assessment guidance is one example of the process used in the 

Superfund program to develop guidance. In that case, Superfund identified the emerging science, 

developed an EPA workgroup to evaluate the available science and its application within the 

Superfund program, released the draft guidance document for public comment, and conducted an 

external peer review. The guidance document (USEPA, 2001a) provides program-specific 

information regarding the conduct of probabilistic risk assessments and supplements the earlier 

policy on this issue (USEPA, 1997c). In addition, EPA has developed training courses on the 

application of this methodology within the Superfund program. To date, probabilistic risk 

assessment methods have been used or are being developed at several sites to evaluate cancer 

risks and non-cancer health hazards (USEPA, 2000m, 2003f). 

The risk assessment processes used at individual Superfund sites have evolved over time 

based on new science and EPA’s understanding of new potential exposure pathways. For 

example, in the early days of the program, dermal exposure was not fully evaluated based on a 

lack of dermal exposure information; this guidance was updated including an external peer 

review (USEPA, 2001g). The Superfund program has updated other guidance documents, 

including external peer review where appropriate to address the current scientific knowledge. 
Examination of EPA Risk Assessment Principles and Practices Page 105 

5.3	 Default Assumptions and Site-Specific Information 

5.3.1	 What Is the Role of Default Assumptions and Site-Specific Information in 

Superfund Risk Assessments? 

Risk assessments incorporate both default assumptions and site-specific information. The 

supplemental guidance document, “Standard Default Exposure Factors” (OSWER Directive 

9285.6-03, March 25, 1991), presents the Superfund program’s default exposure factors for 

calculating RME exposure estimates (USEPA, 1991c). This guidance was developed in response 

to requests that EPA make Superfund risk assessments more transparent and their assumptions 

more consistent. However, the guidance clearly states that the defaults should be used where 

“there is a lack of site-specific data or consensus on which parameter to choose, given a range of 

possibilities.” 

The table in section 5.1 presented examples of default exposure values and the percentile 

of the population the values represent, as well as the peer reviewed studies supporting these 

assumptions. Again, the RME approach uses default values designed to estimate the exposure of 

a high-end individual in the 90th percentile of exposure or above (USEPA, 1992a). Consistent 

with this guidance, relevant default assumptions for various activity levels and age groups are 

used for drinking water consumption rates, soil ingestion rates, residence times, body weight, and 

inhalation rates. The table illustrates the range of percentiles — some defaults included the 50th 

percentile (e.g., body weight), 80th, 90th, and 95th percentiles. 

Although the Superfund program routinely uses default assumptions to assess the risk to 

the RME individual at many sites, the characteristics of the surrounding population change from 

site to site. For example, the distributions of individual residence times will vary depending on 

whether the site is located in a rural or an urban area. Individuals in rural communities are likely 

to have longer residence times than individuals in urban communities. Thus, a default value of 

30 years may fall at the 80th percentile for farmers but above the 95th percentile for renters in an 

urban setting. The extent to which a single default value will impact the final exposure estimate 

depends on the values and variabilities of all the parameters used to estimate exposure. The goal 

is to estimate an individual exposure that actually occurs and is above the 90th percentile. In 

some cases, use of default assumptions may produce an estimate near the 90th percentile; in 

others, the estimate may be higher in the range. 

In general, Superfund’s default factors are designed to be reasonably protective of the 

majority of the exposed population. The assumptions used in Superfund’s risk assessments are 

consistent with the 90th percentile or above and the Agency’s exposure assessment guidelines 

(USEPA, 1992a). Default exposure factors used to assess the RME are a mix of average and 

high-end estimates (see table 5-1). The use of these default exposure assumptions does not 
Page 106 Examination of EPA Risk Assessment Principles and Practices 

automatically result in an overestimation of exposures. These following sections illustrate this 

with several examples. 

5.3.2 What Is the Drinking Water Consumption Rate? 

EPA uses site-specific data on drinking water consumption patterns that have met 

appropriate review criteria at individual sites, but these site-specific studies are generally limited 

or absent. In these cases, to evaluate exposures to contaminants in drinking water, the Superfund 

program applies a default ingestion rate of 2 liters per day for the RME individual (USEPA, 

1989a, 1991c, 2000k). The 2 liters per day assumption represents a reasonable default 

assumption. The results of the 1994 - 1996 Continuing Survey of Food Intake by Individuals 

(CSFII) analysis indicate that the arithmetic mean, 75th, and 90th percentile values for adults 20 

years and older are 1.1, 1.5, and 2.2 liters per day, respectively (USEPA, 2000k). The 2 liters per 

day value represents the 86th percentile for adults. 

In certain parts of the United States — or, during summer seasons, throughout the country 

— people may drink 4 to 4.5 liters per day (USEPA, 1997d). Similarly, a report from OW on the 

analysis of the CSFII data shows that the 99th percentile is 4.2 liters per day (the report provides 

no maximum) (USEPA, 2000k). In some circumstances, then, this default value may 

underestimate cancer risks and non-cancer health hazards for the reasonable high-end exposure 

scenario. 

5.3.3 What Is the Inhalation Rate? 

The 1989 RAGS Part A describes 20 cubic meters per day as an average inhalation rate 

and 30 cubic meters per day as an upper bound (USEPA, 1989a, 1997d). The inhalation rate 

varies depending upon the activity levels of the exposed individual (USEPA, 1997d). The 

default value of 20 cubic meters per day reflects a typical mix of activity levels; however, more 

accurate, site-specific activity level data may be used when available. The default value is 

consistent with the recommendations from the International Commission on Radiological 

Protection (ICRP) data set for the “reference man” (ICRP, 1981; cited in USEPA, 1997d). The 

ICRP used values of 20 to 23 cubic meters per day. These values assumed 16 hours of light 

activity and 8 hours of resting. Daily inhalation rates for individuals performing activities at 

levels other than resting or light activities are not presented. Thus, the values or the exposure 

period may need to be adjusted for individuals (e.g., construction workers) with moderate and/or 

heavy activity levels where inhalation rates would be higher. 

5.3.4 What Is the Exposure Duration for Residences? 

When site-specific data are not available or appropriate, the default residential exposure 

durations are 30 years. This value represents the 90th percentile for home owners and the 90th to 
Examination of EPA Risk Assessment Principles and Practices Page 107 

95th percentile for rural populations (USEPA, 1991c, 1997d). It is not reflective of the residence 

times for farms, where 30 years represent an 80th percentile for this specific population (USEPA, 

1997d). 

Some site-specific exposures may continue even if the individual changes residences, 

provided that he or she remains in the general area of the site. For example, Connelly et al. 

reported that individuals are willing to travel up to 34 miles to continue fishing in a favorite area 

(Connelly et al., 1992). At an existing Superfund site where more than 100 river miles have been 

contaminated, it is thus possible for an angler to change his residence within the surrounding 

counties and continue to fish from the impacted river. An assumption of an exposure duration 

based only on residence time in this instance would underestimate the risks to the angler. 

5.3.5 What Is the Exposure Duration for Workers? 

It is Superfund practice to assume that workers are generally exposed for a duration of 25 

years, unless site-specific data are available (USEPA, 1991c). The exposure duration is assumed 

to be equivalent to job tenure for receptors in an occupational scenario. An analysis of Bureau of 

Labor Statistics data shows that the projected job tenure varies from a few months to 50 years, 

depending on the specific occupational category (Burmaster, 2000). 

More specifically, the projected 95th percentile values for job tenure for men and women 

in the manufacturing sector are 25 years and 19 years, respectively. The projected 95th percentile 

job tenure values for workers in the transportation/utility and wholesale sectors are only 

somewhat less than the values for manufacturing workers — 22 years and 18 years for men and 

women, respectively. Values are lower for other non-industrial sectors — approximately 13 

years for workers in the finance and service sectors, and 7 years for retail workers. Note that the 

95th percentile is within the range for the high-end distribution, which is defined as that part of 

the exposure distribution that is between the 90th and 99.9th percentiles (USEPA, 1997d). 

Thus, the 25-year default value is an estimate of the 95th percentile exposure duration for 

workers across a wide spectrum of industrial and commercial sectors. This value may be an 

overestimate or underestimate of the actual exposure duration, depending on the particular 

circumstances of the employment, e.g., for future workers on a cleaned-up Superfund site. EPA 

supports the use of alternative exposure durations, if they are based on adequate data on job 

tenure and the anticipated industrial/commercial site use. For example, in the evaluation of a 

construction worker scenario, the exposure duration is typically 1 to 2 years, representing the 

time an individual is actually engaged in construction activities on the site (USEPA, 2002k). 

As mentioned above, EPA also typically evaluates CTE scenarios, which is consistent 

with EPA’s guidance on risk characterization (USEPA, 1995a, 2000a). For example, 6.6 years is 

often used as a CTE duration for workers. This value is based on the median occupational tenure 
Page 108 Examination of EPA Risk Assessment Principles and Practices 

of all workers in the United States (USEPA, 1997d). Evaluation of CTE is usually limited to 

comprehensive baseline risk assessments, whereas 25 years is the default exposure duration for 

screening-level assessments in EPA’s waste programs. 

5.3.6 What Is the Ingestion Rate for Construction Workers? 

Comments submitted implied that the Superfund program continues to use an ingestion 

rate of 480 milligrams per day for construction workers. This ingestion rate was updated in 2000 

with EPA’s externally peer reviewed soil screening level guidance to 330 milligrams per day 

(USEPA, 2002k). 

5.3.7 What Are the Incidental Soil Ingestion Rates for Children? 

A common question asked of EPA is why Superfund risk assessments evaluate “dirteating kids”: Why should Superfund sites be cleaned up to levels such that children can safely 

“eat” the soil there? Actually, EPA does not typically assume that children are eating the dirt; 

rather, EPA assumes that they are exposed to contaminants through the course of normal 

activities of play on the ground, exposure to dust in the home, and incidentally through mouthing 

behavior (USEPA 1996b, 2001d). 

It is commonly observed that young children suck their thumbs or put toys and other 

objects in their mouths. This behavior occurs especially among children from 1 to 3 years old 

(Behrman and Vaughan, 1983; Charney et al., 1980). This “hand-to-mouth” exposure is well 

documented in the scientific literature for children under 6, and is especially prevalent among 

children 11⁄2 to 3 years old, a critical period for brain development. This time period is of special 

concern regarding potential exposure, since children may be at special risk of exposure to 

specific chemicals, e.g., lead (ATSDR, 1991). Superfund experience has taught us that children 

do incur exposures to contaminated soil, as is evident at lead-contaminated sites in which 

elevated blood levels occur in children residing at those sites (USEPA 1996b, 2001d). 

Scientists agree that because of this behavior, children may incidentally or accidentally 

take in soil and dust (Calabrese et al., 1989; Davis et al., 1990; Van Wijnen et al., 1990). Where 

children are likely to be exposed to contaminated soils (in residential areas, for example), it is 

appropriate for EPA to evaluate potential risks and set cleanup levels that will protect children 

for this widely recognized pathway of exposure, especially during this sensitive developmental 

period in the child’s lifetime. 

The basis of EPA’s default soil ingestion rate is generally a point of contention. EPA has 

developed soil ingestion rates that are used as “default exposure assumptions” for adults and 

children. For young children (6 years or younger), the Superfund program default value is 200 

milligrams of soil and dust ingested per day (USEPA, 1991c, 1996a). EPA’s risk estimates 
Examination of EPA Risk Assessment Principles and Practices Page 109 

address the “incidental” ingestion that might occur when a child puts a hand or toy in his or her 

mouth, or eats food that has touched a dusty surface. Although this default assumption is often 

presented as an overly conservative value, the amount (200 milligrams per day) represents a 

small amount of soil ingested. It is less than 1/100 of an ounce (or one-fifth of the contents of a 

single-serving packet of sugar, e.g., a small package of Equal sweetener) a day. This peer 

reviewed value is applied in estimates of RME exposures (USEPA, 1989a, 1991c, 1997d). 

In Superfund risk assessments, this soil ingestion rate for young children is combined 

with site-specific assumptions about exposure frequency (days per year) to estimate an average 

intake over the 6-year exposure period. Exposure frequency varies depending on site-specific 

current and future land uses. Soil ingestion studies report daily averages; the amount of soil 

ingested cannot be prorated on an hourly basis. Also, soil ingestion is episodic in nature and 

dependent upon a child’s activity patterns, so prorating by time is not always appropriate. This is 

a common misapplication of soil ingestion rates in risk assessment. 

Some children deliberately eat soil and other non-food items (a behavior known as pica). 

Pica behavior has been identified in children at rates of up to 5,000 milligrams per day (ATSDR, 

1996, 2000; Calabrese et al., 1991). The Agency for Toxic Substances and Disease Registry uses 

this pica ingestion rate when calculating Environmental Media Evaluation Guides, which are 

used to select contaminants of concern at hazardous waste sites (ATSDR, 1996). EPA itself does 

not routinely address this form of exposure unless site-specific information is available. The 

default soil ingestion rate of 200 milligrams per day applied in Superfund risk assessments is 

intended to ensure reasonable protection of children in cases where they are likely to become 

exposed to contaminated soils and dust associated with a Superfund site. 

5.4 Site-Specific Information 

5.4.1 When Does EPA Use Site-Specific Information? 

Again, site-specific information is considered in risk assessment where available. EPA 

risk assessment guidance supports the use of site-specific data, where feasible and appropriate 

according to a careful evaluation of the information including the specifics of the study design, 

numbers of individuals evaluated, representativeness of the exposures, etc. Examples of useful 

information are location of residences (receptor location), current and future land uses, behavior 

patterns and preferences (drinking water use, fish preparation), unique exposure pathways 

(hunting of specific game), and residence time within specific geographic areas. 

The nature of the data submitted by stakeholders varies from anecdotal information to 

knowledge about creel surveys or other surveys conducted by universities or other groups. The 

nature and amount of data submitted will vary based on the site, ranging from no submission to 
Page 110 Examination of EPA Risk Assessment Principles and Practices 

data on specific work practices at a facility. When such information is submitted, EPA critically 

evaluates it to determine its use in the risk assessment portion of the remedial investigation (RI). 

The most accurate way to characterize potential site-specific exposures to populations 

around Superfund sites would be to conduct a detailed census at each site considering both 

current and future land uses. Theoretically, this would involve interviewing all potentially 

exposed individuals regarding their lifestyles, daily patterns, water usage, cleanliness, 

consumption of local fish and game and procedures, working locations, and exposure conditions, 

while collecting environmental samples (e.g., indoor air, water, soil, food, and possibly more). 

Although site-specific data are collected on environmental media (e.g., soil, groundwater, air, 

etc.) as appropriate during the RI, such collection has significant limitations. The three almost 

insurmountable difficulties are time, expense, and a major intrusion on privacy. 

Discussed below are two examples of exposure parameters for which EPA has collected 

and applied site-specific data in risk assessment: bioavailability and fish consumption rates. 

5.4.2 How Is Bioavailability Addressed? 

The Agency addresses bioavailability by using default values and, in some cases, 

developing site-specific values supported by laboratory studies. In the absence of reliable 

information, RAGS Part A recommends assuming that the relative absorption efficiency between 

food or soil and water is 1. In other words, the assumption is that the bioavailability of the 

contaminant on the site, regardless of exposure medium, is the same as the bioavailability in the 

toxicity study used to derive the RfD or CSF. Such a default value is used to ensure that we do 

not underestimate risk in the face of uncertainty. The bioavailability of a compound in the 

exposure medium of concern at the site may actually be greater than in the exposure medium 

used in the critical toxicity study that formed the basis of the RfD or CSF. If this is the case, 

assuming a relative bioavailability of 1 for the medium of concern may result in an underestimate 

of risk at the site, depending on how all the other parameters are evaluated. 

Appendix A of RAGS Part A (USEPA, 1989a) specifically addresses the consideration of 

site-specific bioavailability information in human health risk assessments. This information is 

referred to as “adjustments for absorption efficiency.” In particular, Appendix A states that sitespecific bioavailability adjustments may be appropriate: 

Adjustments also may be necessary for different absorption efficiencies depending on the medium of 

exposure (e.g., contaminants ingested with food or soil might be less completely absorbed than 

contaminants ingested with water). 

If the medium of exposure in the site exposure assessment differs from the medium of exposure assumed by 

the tox icity value ( e.g., RfD values u sually are base d on or ha ve be en ad justed to reflec t expo sure via 

drinking water, while the site medium of concern may be soil), an absorption adjustment may, on occasion, 

be ap pro priate . For e xam ple, a su bstan ce m ight be mor e com pletely a bso rbed followin g exp osure to 
Examination of EPA Risk Assessment Principles and Practices Page 111 

contaminated drinking water than following exposure to contaminated food or soil (e.g., if the substance 

does not desorb from soil in the gastrointestinal tract). 

For example, juvenile swine have been used as a model to derive site-specific 

bioavailability adjustments at several mining and smelter sites across the country (Casteel et al., 

1997; USEPA, 1999g). These adjustments have been primarily limited to lead and arsenic. 

Default values have been developed for some metals. For example, lead risks in children 

are typically assessed by predicting blood lead levels using the Integrated Exposure Uptake 

Biokinetic Model for Lead in Children (IEUBK) developed by EPA (USEPA, 2001d). The 

IEUBK model assigns default absolute bioavailability factors to all lead exposure media. The 

default values for air, water, and soil are 32%, 50%, and 30%, respectively (USEPA, 1999g, 

2001d). For nonresidential scenarios, EPA developed the Adult Lead Methodology, which 

assumes the absolute bioavailability of lead in soil is 12% (USEPA, 1996b). 

EPA finds that the site-specific application of quantitative bioavailability adjustments in 

risk assessments is not frequently supported by available scientific data. This occurs principally 

because of a lack of validated data and models to assess bioavailability for common site 

contaminants. 

5.4.3 How Are Fish Ingestion Rates Evaluated? 

Fish ingestion is evaluated at many Superfund sites with contaminated river systems, 

streams, and lakes. These contaminated sites range in size from small creeks to large river 

systems covering hundreds of miles. The contaminants of concern found at these sites include 

PCBs, dioxins and furans, mercury, pesticides, and other chemicals identified as persistent 

organic pollutants. State health departments have established fish consumption advisories at 

many of these water bodies. 

The 1989 RAGS Part A guidance (USEPA, 1989a) recognized the importance of 

evaluating the potential cancer risks and non-cancer health hazards associated with consuming 

contaminated fish associated with Superfund sites. Page 6-6 of RAGS Part A specifically states: 

... Be sure to include potentially exposed distant populations, such as public water supply consumers and 

distant c onsu mers of fish or sh ellfish or a gricultur al pro duc ts from th e site are a. ... 

Further, on page 6-7, the guidance states: 

... Identify any site-specific population characteristics that might influence exposure. For example, if the 

site is located near major commercial or recreational fisheries or shell fisheries, the potentially exposed 

pop ulation is like ly to eat m ore lo cally-cau ght fish and shellfish than pop ulations locate d inlan d. ... 
Page 112 Examination of EPA Risk Assessment Principles and Practices 

A variety of data sources are available to evaluate fish ingestion rates. Fish ingestion 

studies can be either creel surveys, in which anglers are interviewed in person while fishing, or 

mail surveys, in which anglers (often identified as people with fishing licenses) are sent 

questionnaires in the mail (reviewed in USEPA, 1992c). Creel surveys typically involve 

interviews with anglers at the dockside requesting information about their fishing activities (fish 

preference, consumption rates, cooking methods, age, gender, frequency of fishing the specific 

water body, etc.). This survey method can provide information on both licensed and unlicensed 

anglers, depending upon who is interviewed. Mail surveys, meanwhile, typically involve sending 

questionnaires to licensed anglers requesting information on fishing practices; preferred rivers, 

lakes, or streams; fish consumption; and other information. If mailing addresses are obtained 

from lists of licensed anglers, unlicenced anglers are not represented. A third type of survey, 

diary surveys — in which participants are asked to record the frequency of fish ingestion, the 

types of fish eaten, and the meal size — require more effort from survey participants, but are 

generally assumed to yield more accurate results because they minimize the potential recall bias 

found in the other survey methods. 

Any of these methods may be appropriate to evaluate the impact of Superfund site 

contamination on individuals consuming fish from contaminated water bodies impacted by a site. 

The approaches used in assessing fish consumption rates at sites vary based on the availability of 

creel surveys, the size of the impacted water body (e.g., a creek as opposed to a large river 

system), the available information on fishing practices within the specific water body, and the 

time and resources necessary to conduct a site-specific survey of fish consumption patterns. 

The risk assessments include an evaluation of the RME and CTE individuals. In the 

absence of site-specific fish consumption rates, the default values from the 1997 Exposure 

Factors Handbook (USEPA, 1997d) may be used. The handbook makes a number of 

recommendations based on 95th and 50th percentile fish ingestion rates for specific water bodies, 

geographic locations, and fish species (USEPA, 1997d). Table 5-2, below, provides some of 

those recommendations. 
Examination of EPA Risk Assessment Principles and Practices Page 113 

Table 5-2: Comparison of Default Fish Ingestion Rates 

Ingestion 

Rate 

Default (grams/day) Percentile Data Source 

1997 EPA Exposure Factors 

Handbook 

General population

 Long-term intake

 distribution 63 95th USEPA, 1997d, based on 

TRI (Javitz, 1980; Ruffle et 

al., 1994) 

Total fish 20.1 50th USEPA Analysis of CSFII, 

 Marine fish 14.1 50th 1989–1991 

Freshwater/estuarine 6.0 50th 

1997 EPA Exposure Factors 

Handbook (recreational 

marine anglers)

 Atlantic 18.0 95th USEPA, 1997d; NMFS, 

5.6 50th 1993 

Gulf 26.0 95th 

7.2 50th 

Pacific 6.8 95th 

2.0 50th 

1997 EPA Exposure Factors 

Handbook (freshwater) 

Maine 13 95th Ebert et al., 1993 

5 50th 

New York 18 95th Connelly et al., 1996 

5 50th 

Michigan 39 95th West et al., 1989 

12 50th 

Michigan — 95th West et al., 1993 

17 50th 
Page 114 Examination of EPA Risk Assessment Principles and Practices 

Species-specific intake information may be derived from a state-wide angler survey 

(Connelly et al., 1992) to calculate the concentration of PCBs ingested in fish. For example, at a 

PCB site, EPA determined that each species of fish has a characteristic PCB concentration and 

the average concentration an angler consumes will, in part, be based on the relative percentages 

of the different fish species consumed. This analysis yields both central tendency and RME 

estimates for each of the parameters needed to calculate the cancer risks and non-cancer health 

hazards. A site-specific analysis using this database found that, for adults, the central tendency 

fish ingestion rate was 4.0 grams per day, or about six half-pound meals per year; the RME fish 

ingestion rate was 31.9 grams per day, or about 51 half-pound meals per year for adults. Fish 

ingestion rates for adolescents and young children were reduced based on the ratio of adolescent 

or child body weight to that of an adult. 

Other risk assessments have used studies from the published literature that are 

representative of the relevant site-specific consumption patterns. Risk assessors review the 

literature to determine how well it represents the site they are evaluating. In determining the 

literature’s appropriate use, they may evaluate the study design, hypotheses being tested, number 

of individuals interviewed within the survey, rates of consumption, and populations studied. 

In several other examples, site-specific surveys have been conducted to evaluate the 

consumption patterns for specific populations that the published surveys do not capture. These 

surveys found considerably higher consumption rates among these populations than if the 

standard default assumptions from the 1997 Exposure Factors Handbook were used (USEPA, 

1997d). For example, a 31⁄2-year site-specific creel survey (Toy et al., 1996) included 

information on whether or not adults harvested fish and shellfish from Puget Sound. The survey 

included 190 adults and 69 children between the ages of 0 and 6. The study found that tribal 

seafood consumption rates were considerably higher than Exposure Factors Handbook values. 

Among the Squaxin, the average consumption rate was 72.8 grams per day and the 90th percentile 

ingestion rate was 201.6 grams per day. Among the Tulalips, the average consumption rate was 

72.7 grams per day and the 90th percentile was 192.3 grams per day. Other site-specific 

consumption surveys found similar differences in consumption rates (APEN, 1998; USEPA, 

2001h; Sechena et al., 2003). 

In cases where EPA has conducted individual surveys to identify fish consumption rates, 

EPA has found it important to include the community in the process (USEPA, 1999f). EPA and 

other agencies (both private and governmental) have spent considerable resources and time to 

plan and implement these studies. The surveys (APEN, 1998; USEPA 2001h; Sechena et al., 

2003) were all conducted using one-on-one interviews, as opposed to creel or mail surveys. The 

people conducting the interviews were always specially trained members of the ethnic group or 

community being surveyed. (EPA realized early on that creel and mail surveys do not necessarily 

work for tribes and other such groups. One-on-one interviews involving the affected 

communities are more appropriate.) 
Examination of EPA Risk Assessment Principles and Practices Page 115 

The determination of the need to conduct a creel survey or use site-specific information is 

dependent on a number of site-specific factors. These include available information on 

consumption patterns, whether an existing consumption advisory may bias responses when 

surveys are conducted, time, and resources. 

5.5 Other Factors in Superfund Assessments 

5.5.1 What Is the Role of Stakeholders in Superfund Assessments? 

EPA works with members of the community to understand their concerns at the national 

and site-specific levels. In 1996, EPA conducted meetings with the communities to discuss the 

Superfund process and their concerns (USEPA, 1996c). Under the 1996 RAGS Reform process, 

EPA met with community members at several meetings to discuss their concerns and identify 

priorities for new guidance documents. At these meetings, the communities identified ideas for 

improving communication between the regulators and the community. This resulted in guidance 

for involving communities in the risk assessment process (USEPA, 1999f). 

Within the Superfund program, EPA conducts extensive outreach with stakeholders at the 

site. As described in the guidance on community involvement (USEPA, 1999f, 2000n), at each 

site the remedial project manager, the community involvement coordinator, and other members 

of the team including the risk assessor regularly communicate with the members of the 

community, including the potentially responsible parties. Each site has a Community 

Involvement Plan, developed following interviews with the community, that outlines 

mechanisms by which the community will be involved in the process. 

EPA guidance and educational materials help illustrate the ways that citizens can be 

involved in the risk assessment process (USEPA, 1999f, h). For example: Community-specific 

information on fishing preferences helped to identify exposure areas for sampling and fish 

species consumed by people who fish in a contaminated bay. Information from farmers on 

pesticide applications helped EPA determine why certain contaminants were present in an 

aquifer. Discussions with farmers about certain harvesting practices helped EPA refine exposure 

models and assumptions at another site (USEPA, 1999h). 

EPA uses a range of communication tools to include the community in the Superfund 

process. These include newsletters, fact sheets, site-specific home pages, public meetings, public 

availability sessions, and 1-800- numbers to contact EPA staff. EPA strives to communicate 

information about the RI, the results of the risk assessment, proposed actions at the site, and the 

proposed and final decisions for remedial actions. The Record of Decision (ROD) includes a 

responsiveness summary that addresses comments including those from the community. During 

the period of the remedial action, communication with the community continues, including 

updates during the 5-year review process. 
Page 116 Examination of EPA Risk Assessment Principles and Practices 

EPA conducts peer reviews of site-specific risk assessments following appropriate peer 

review guidance (USEPA, 2000b). 

Consistent with EPA Superfund goals of improving the transparency of the process, the 

methods for summarizing risk information are found in the RAGS Part D (USEPA, 2001b). 

Superfund continues to update guidance documents to improve the transparency of risk 

information. 

5.5.2	 How Are Superfund Assessments Factored in Superfund Risk Management 

Decisions? 

Risk assessment is one of several tools used to inform risk management decisions. Risk 

managers weigh a number of factors, including uncertainties in exposure and risk estimates, 

when developing health and environmental protective decisions. EPA considers a variety of 

alternatives to protect human health and the environment at sites and evaluates them by 

considering long-term effectiveness, use of treatment, short-term effectiveness, implementability, 

and cost. EPA then proposes a protective, cost-effective, ARAR-compliant (that is, compliant 

with the Applicable or Relevant and Appropriate Requirements (ARAR)) remedy, which it may 

modify based on state and public comments (see also CERCLA § 121, 42 U.S.C. § 9621 and 40 

CFR § 300.430[e][9]) (USEPA, 1988b, 1999a, f, 2000n; NCP 1990a, b). 

Following the completion of the RI during which the risk assessment is conducted, EPA 

develops a feasibility study (FS) that evaluates remedial actions at the site (USEPA, 1988b). The 

FS evaluates the risks in the absence of remedial actions or institutional controls. This provides a 

baseline for comparison with other remedial alternatives. The FS includes the development of 

Remedial Action Objectives, including Preliminary Remediation Goals (PRGs) that are 

developed based on the RME exposure assumptions used in the risk assessment. The PRGs 

provide concentration levels that are protective of the RME individual who is currently exposed 

or may be exposed in the future. 

During the FS, remedial alternatives are developed to achieve the program goals through 

a variety of different methods, generally including containment and treatment alternatives. The 

alternatives reflect the scope and complexity of the site problem. The Superfund program 

evaluates these alternatives using nine criteria described by the NCP. The criteria address 

protectiveness, effectiveness, implementability, and acceptability issues. The criteria were 

derived from remedy selection criteria provided by Congress in SARA 121 (SARA, 1986). The 

detailed analysis consists of an assessment of the individual alternatives against each of the nine 

evaluation criteria and a comparative analysis focusing upon the relative performance of each 

alternative against those criteria. The nine criteria are presented in table 5-3, below. In addition 

to viable remedial alternatives, EPA evaluates a no-action remedial alternative at all sites. The 

no-action alternative provides a baseline for comparison of the various alternatives that are 
Examination of EPA Risk Assessment Principles and Practices Page 117 

appropriate for a specific site. All of this information is provided in a Proposed Plan, which is 

released with the RI/FS for public review and comment. 

EPA provides opportunities for public review of this information. A public meeting is 

held to discuss the proposed remedial alternatives and to obtain comments. Public comments are 

addressed at the meeting and in the Response to Comments that is developed as part of the ROD. 

The ROD identifies remedial actions that have been selected for the site. 

Following the ROD, EPA begins the remedial design process and the implementation of 

construction. Depending on the nature of the remedial actions and the amount of time required 

to complete the construction, EPA may conduct 5-year reviews to determine the protectiveness of 

the remedy (USEPA, 2001i). Throughout this process, information is shared with the affected 

community regarding the progress of the remedial actions. 
Page 118 Examination of EPA Risk Assessment Principles and Practices 

Table 5-3: Nine Evaluation Criteria for Superfund Remedial Alternatives 

Threshold Criteria 

Overall protection of human health and the environment determines whether an alternative 

eliminates, reduces, or controls threats to public health and the environment through 

institutional controls, engineering controls, or treatment. 

Compliance with ARARs evaluates whether the alternative meets federal and state 

environmental statutes, regulations, and other requirements that pertain to the site, or whether 

a waiver is justified. 

Primary Balancing Criteria 

Long-term effectiveness and permanence considers the ability of an alternative to maintain 

protection of human health and the environment over time. 

Reduction of toxicity, mobility, or volume of contaminants through treatment evaluates an 

alternative’s use of treatment to reduce the harmful effects of principal contaminants, their 

ability to move in the environment, and the amount of contamination present. 

Short-term effectiveness considers the length of time needed to implement an alternative and 

the risks the alternative poses to workers, residents, and the environment during 

implementation. 

Implementability considers the technical and administrative feasibility of implementing the 

alternative, including factors such as the relative availability of goods and services. 

Cost includes estimated capital and annual operation and maintenance costs, as well as present 

worth cost. Present worth cost is the total cost of an alternative over time in terms of today’s 

dollar value. Cost estimates are expected to be accurate within a range of +50% to -30%. 

Modifying Criteria 

State acceptance considers whether the state agrees with the EPA’s analyses and 

recommendations, as described in the RI/FS and Proposed Plan. 

Community acceptance considers whether the local community agrees with EPA’s analyses 

and preferred alternative. Comments received on the Proposed Plan are an important indicator 

of community acceptance. 
Examination of EPA Risk Assessment Principles and Practices Page 119 

5.6	 Superfund Site-Specific Conclusions 

In general, the information applied in EPA Superfund risk assessments and the level of 

protectiveness are considered in the context of the regulatory program. From the previous 

discussion, the following conclusions are highlighted. 

a)	 Reasonable maximum exposures. For the Superfund program, EPA bases 

decisions on current and future risks associated with reasonable high-end 

exposures or RME, not only the average exposures. By definition, the RME is 

within a realistic range of exposures, not a “worst-case” estimate. 

b)	 Default assumptions. Risk estimates are based on a combination of site-specific 

information, when available and appropriate, and general default exposure factors 

developed for Superfund baseline risk assessments. The default factors are based 

on published scientific data and are designed to be reasonably protective of the 

majority of those exposed. The use of defaults also streamlines risk assessments, 

reduces unwarranted variability in the exposure assumptions used to characterize 

exposures, and provides reasonable risk estimates where there is a lack of 

site-specific data or consensus on which value to choose. For some exposure 

parameters, default values are used because the costs of collecting data for 

specific behaviors in a population may be prohibitive. 

c)	 Site-specific information. Information submitted to the EPA regions or collected 

by EPA is critically evaluated to determine how to apply it in the risk assessment. 

Anecdotal information is valuable but not sufficient for quantifying exposures. It 

may, however, be useful to consider such information when developing a 

conceptual site model and characterizing uncertainties. In some cases, sitespecific data would be preferable to using default exposure data that represent the 

majority of the population at the high end of the distribution for specific activity 

patterns. However, the costs of collecting data for specific behaviors in a 

population may be prohibitive. For example, a site-specific creel survey may cost 

$100,000 or more and take several years to conduct (Toy et al., 1996; USEPA 

2001h; Sechena et al., 2003). 

Specialized studies may reduce uncertainties in a site-specific risk assessment; 

however, the region considers the sensitivity of the parameter in question and the 

added value of the information in light of uncertainties of other parameters, 

including the toxicity information. In some cases, a research study of exposure 

will not significantly reduce the uncertainty of the risk assessment when the 

toxicity information is very limited (USEPA, 1992a). 
Page 120 Examination of EPA Risk Assessment Principles and Practices 

d) Peer review. Consistent with Agency guidance on peer review, external peer 

reviews are conducted for specific risk assessments and guidance documents. 

e) Probabilistic risk assessment. Superfund has developed specific guidance for 

conducting site-specific probabilistic risk assessments, and this guidance was 

externally peer reviewed. At the present time, site-specific probabilistic risk 

assessments have been conducted at several sites to examine exposure 

assessments. 

f) Guidance documents. The Superfund program continues to update guidance to 

include new science where appropriate. The guidance documents are made 

available for public review and external peer review. 

g) Transparency and clarity in risk assessments. Superfund developed specific 

guidance for improving the clarity and transparency of risk assessments. This 

guidance covers community involvement, standardization of reporting of risk 

information, and standardization of ROD information (USEPA, 1999f, 2000n). 

EPA’s risk assessment practices and policies apply a mix of site-specific information and 

science-based default assumptions that are consistent with the goals of the Superfund program. 

Site-specific considerations are addressed at the EPA site-specific level. In turn, the regions 

inform EPA program offices of the research needed to reduce uncertainties in risk assessments. 

There will always be a need for additional research on exposure and toxicity parameters to 

improve understanding of how people contact their environments, chemical toxicity, and 

chemical fate and transport. 

5.7 Chemical Mixtures Risk Assessment Methods and Practice 

5.7.1 What Are the Issues Regarding Chemical Mixture Site Assessments? 

EPA published guidance on the health risk assessment of chemical mixtures (USEPA, 

1986b; 2000h) and its use in site assessments (USEPA, 1989a; 1999a). As part of the 

implementation of the 1996 Food Quality Protection Act, EPA also developed guidance for 

conducting cumulative risk assessments of chemicals that appear to act by a common mechanism 

of toxicity (USEPA, 2002f). EPA (USEPA, 2000h) guides the risk assessor through an 

examination of data relevant to a mixture of concern to facilitate selection of an appropriate 

mixture risk assessment method. Suggested methods include both whole-mixture approaches 

and component-based algorithms that include interaction data. A comment regarding EPA’s 

practice is that the Agency “assumes the toxicity of a chemical mixture is equal to the sum of the 

toxicity of each individual chemical, regardless of the toxicity type, competition or antagonism 

among chemicals.” This comment apparently assumes that: 
Examination of EPA Risk Assessment Principles and Practices Page 121 

a)	 EPA uses only models of mixture toxicity based on dose addition, 

b)	 EPA inappropriately applies dose addition by ignoring the requirement that the 

mixture’s chemical components share a common toxic mode of action (MOA), 

c)	 EPA ignores interactions data that may reduce the toxicity of a mixture through 

competition or antagonism, and 

d)	 EPA never determines the toxicity of a whole mixture directly. 

These assumptions do not reflect current guidance, as shown by the Agency’s mixtures 

guidelines (USEPA, 1986b) and supplementary guidance (USEPA, 2000h). These documents 

offer detailed descriptions of a variety of component-based and whole-mixture methods far 

beyond dose addition, provide instructions on appropriate application of the methods, and 

recommend consideration of interaction data whenever possible. Further, program guidance 

(USEPA, 1989a) specifies the use of component-based approaches to evaluate chemical 

mixtures. The following sections explain the theory behind these procedures and discuss current 

practice. 

5.7.2	 What Are Dose Addition and Response Addition? 

To address low exposure levels when no interaction information or whole-mixture 

toxicity data are available, it is EPA’s practice to use response addition and dose addition as the 

recommended default methods when the component chemicals in a mixture show dissimilar 

toxicity and similar toxicity, respectively (USEPA, 2000h). Dose addition and response addition 

are fundamentally different methods, based on different assumptions about toxicity. Because 

both methods are relatively easy to apply and use single chemical toxicity and exposure 

information, they are the methods most commonly used in site-specific assessments (USEPA, 

1989a). The two additivity assumptions are further described as follows: 

a)	 Dose addition sums the doses of the components in a mixture after they have been 

scaled for toxic potency relative to each other. The risk of a toxic effect is 

determined from this summed dose. Dose addition requires the components to 

share a common toxic MOA. 

Superfund site assessments have applied dose addition in the form of a hazard 

index (HI) to evaluate sites for indications of health risk (USEPA, 1989a). The HI 

is calculated as the sum of hazard quotients (HQs) for the chemical components of 

the mixture; thus, it indicates risk, but is not an explicit risk estimate. A HQ is 

typically calculated as the ratio of a chemical’s exposure level to its safe level, 

such that values larger than 1 are of concern. For a group of n chemicals in a 
Page 122 Examination of EPA Risk Assessment Principles and Practices 

mixture with exposure levels (E) and using the RfD as a safe level, the HI is 

calculated as: 

Equation 5.7-1 

The HI is usually calculated for groups of chemicals whose effects are observed in 

a common target organ. As an initial screening step, though, risk assessors 

calculate HIs by combining across health endpoints, contradictory to the 

assumption of common MOA. If the HI across all effects indicates no risk (i.e., is 

less than 1), then the assessment can stop. If risk is indicated (i.e., the HI is 

greater than 1), the risk assessor repeats the analysis by breaking the mixture into 

groups of chemical components with a common target organ and recalculating the 

HI for each smaller group. 

b)	 Response addition first estimates the probabilistic risk of observing a toxic 

response for each chemical component in the mixture. Then the component risks 

are summed to estimate total risk from exposure to the mixture, assuming 

independence of toxic action (i.e., that the toxicity of one chemical in the body 

does not affect the toxicity of another chemical). This can be thought of as an 

organism receiving two (or more) independent insults to the body, so that the risks 

are added under the statistical law of independent events. 

Superfund site assessments calculate total cancer risk by summing the individual 

cancer risks for the carcinogens in the mixture (USEPA, 1989a). For example, the 

mixture risk (Rm) for two chemicals is the sum of the risks for chemical one (r1) 

and chemical two (r2) minus the probability that the toxic event from exposure to 

chemical one would overlap in time with the toxic event from exposure to 

chemical two, as follows: 

Equation 5.7-2 

Because cancer risks are typically very small (i.e., in the 10-4 to 10-6 range), the 

amount of risk subtracted in Equation 5.7-2 (e.g., an amount in the 10-8 to 10-12 

range) is insignificant and is therefore usually ignored. Risks are appropriately 

aggregated for cancers across various target organs because the result is 

interpreted as the risk of any cancer, and the cancers from each chemical 

component are considered to be independent events in the body. 
Examination of EPA Risk Assessment Principles and Practices Page 123 

5.7.3 How Are Dose Addition and Response Addition Applied in Practice? 

To get a general idea of how dose addition and response addition are actually practiced in 

EPA’s site assessment work, a search of EPA’s RODs in the Comprehensive Environmental 

Response, Compensation, and Liability Information System (CERCLIS) database was conducted 

in August 2003 (http://www.epa.gov/superfund/sites/siteinfo.htm), keying on the term “hazard 

index.” From that list, 10 RODs were selected from across the Agency. These RODs were 

examined for tables and text relevant to the application of response addition and the Hazard 

Index (HI). RODs were used as a surrogate for the complete risk assessments since only a 

limited number of risk assessments are available on the internet. It is noted that, as stated in the 

guidance for preparing RODs, the full risk assessment need not be included in the ROD 

document (USEPA, 1999a). The risk assessments are available in the Administrative Record for 

each site and, based on the time limits of this analysis, it was not possible to evaluate each of the 

individual site records. 

To estimate cancer risks using response addition, the procedure is to sum risk estimates 

across all carcinogenic modes of action, assuming independence of action. There is no 

requirement to demonstrate independence empirically; however, when component exposures are 

low, such biological independence is likely. Most of the examples appear to have appropriately 

applied response addition, although in several cases, this conclusion is inferred from the 

available text. In one instance, the technique was not used because there was only one 

carcinogen found at the site. In two other cases, additional information from the risk assessment 

would be needed to determine exactly how the evaluations were completed. 

To apply the HI, the procedure in the guidance is to first perform a screening-level 

analysis, summing Hazard Quotients (HQs) across all target organs. If the HI is below 1 across 

all target organs, then the analysis is complete; this was the case for tap water risks in one 

example. If the HI is above 1 across all target organs, then the analysis is to be repeated by 

summing the HQs for only those mixture components that affect the same target organ; this 

occurred in the three of the examples. The purpose of the re-analysis is to comply with the dose 

addition assumption of a common toxic mode of action. In one case, the HI was shown in the 

ROD to change from greater than 1 in the screening analysis to less than 1 under the re-analysis. 

In two instances, the HI was greater than 1 in the screening analysis, but the ROD did not show 

the re-analysis by target organ; in these cases, however, the individual chemical HQs were much 

larger than 1, making all HI calculations in these cases moot (i.e., no HI below 1 could be 

calculated under any constraints). Finally, in four of the examples, additional information would 

be needed to determine exactly how the evaluations were completed. 

In conclusion, the results from this limited analysis are generally positive, illustrating that 

EPA practice is consistent with its own guidance regarding chemical mixtures risk assessment. 

The clarity of EPA’s risk communications could be improved by transferring additional 
Page 124 Examination of EPA Risk Assessment Principles and Practices 

information from the risk assessment in the Administrative Record for each site to its associated 

ROD. 

5.7.4 Are Interaction Data Used in Mixture Risk Assessment? 

EPA defines interactions as observed effects greater than (synergism) or less than 

(antagonism) those expected under a specified form of additivity. EPA (USEPA, 1986b, 2000h) 

also states that mixture risk assessment should reflect known toxicological interactions. In 

practice, sufficient data are seldom available to model interactions. Data on binary combinations 

of chemicals are most prevalent, so for mixtures of more than two chemicals, the true nature of 

joint toxic action may be speculative at best. For exposures at low doses with low component 

risks, the likelihood of a significant interaction is usually considered to be low. Interaction 

arguments based on saturation of metabolic pathways or competition for cellular sites usually 

imply an increasing interaction effect with dose, so that the importance for most low 

environmental exposures is probably small. 

EPA mixture risk assessments have usually considered the information on interactions in 

a qualitative sense. For example, a Superfund site may receive more scrutiny or its remediation 

may proceed faster if there are indications of potential synergism among the detected chemicals. 

The cleanup goals and the estimated risk would not change, though, because the synergism could 

not be quantified. No standard methods or mathematical models are yet in common use in 

regulatory agencies to incorporate interactions and to serve as defaults. There is one newly 

developed procedure, the interaction-based HI (USEPA, 2000h), that allows for numerical 

adjustments to the HI based on evidence of synergism or antagonism for pairs of chemicals in the 

mixture. This method is currently undergoing verification using laboratory data. 

Despite the difficulties in using interaction information, some regional assessments have 

considered relevant data. An interesting assessment by Region III explicitly evaluated synergy 

and antagonism of multiple metals in soil (USEPA, 2002m). Text from that ROD reads as 

follows: 

EP A (1 994 b; [this reference is in the ROD]) prepared an extensive review of interactions among cadmium, 

lead, and zinc reported in the open literature. Available evidence does not support a change in assumed 

absorption of metals due to interactions among metals following ingestion of contaminated soils. The 

quantitative risk assessment does not alter the estimates of exposure based on co-exposure to cadmium, 

lead, and zinc. Further, available information suggests that any potential impact of cadmium on 

neurotoxicity of lead in young children is largely speculative. No modification to the assessment of lead 

toxicity is thus justified based on co-exposure to cadmium. 

5.7.5 Are Data on Whole Mixtures Used in Risk Assessment? 

Risk assessments based on tests of whole mixtures or on epidemiologic data determine 

combined effects empirically. Examples of these (USEPA, 2003g) include (1) RfDs on 
Examination of EPA Risk Assessment Principles and Practices Page 125 

commercial PCB mixtures (Aroclors 1016 and 1254) based on primate data, and (2) a CSF for 

coke oven emissions based on human occupational exposures. Drinking water disinfection 

byproducts represent a complex mixture for which epidemiologic data suggest potential health 

risks (USEPA, 2003b). An EPA study, called the 4 Lab project, is currently underway to 

toxicologically and chemically characterize this complex mixture to produce data on 

reproductive and developmental effects in rats exposed to concentrations of this complex mixture 

for use in risk assessment (Simmons et al., 2002). 

Whole-mixture studies are routinely used in ecological risk assessments. The Agency has 

developed subchronic toxicity tests for whole aqueous effluents and for contaminated ambient 

waters, sediments, and soils (USEPA, 1989d, 1991d, 1994e). Further, the effects of mixtures in 

aquatic ecosystems are evaluated using bioassessment techniques that are equivalent to 

epidemiology, but more readily employed (USEPA, 1999i). Similar bioassessment methods are 

sometimes used at Superfund sites (USEPA, 1994f). These empirical approaches to assessing 

ecological risks from mixtures are employed in National Pollutant Discharge Elimination System 

permitting and the development of Total Maximum Daily Loads, and are often used in Superfund 

baseline ecological risk assessments. 

5.7.6	 What Other EPA Applications Exist for Mixtures Risk Assessment? 

EPA program offices and regional risk assessors have a great need for both assessment 

information and risk assessment methods to evaluate human health and ecological risks from 

exposure to chemical mixtures. Further, several pieces of legislation exist that require EPA to 

consider the evaluation of chemical mixtures (see below). The Agency conducted some specific 

mixture risk assessments in the past and continues to pursue additional evaluations. Currently, 

an IRIS assessment is being developed for the PAHs (USEPA, 2002n). In addition, EPA used a 

dose additive approach to assess PCBs (USEPA, 1996d), organophosphorus pesticides (USEPA, 

2002g), and dioxins (USEPA, 1989b; 2000i). (See section 4.5 for a discussion of this approach.) 

Examples of other mixtures of interest to the Agency include creosote, particulate matter, 

drinking water disinfection byproducts (DBPs), brominated flame retardants, mixtures of metals, 

and pesticide mixtures in soils. Current Agency interests and supporting legislation include: 

a)	 OW is concerned with contaminant mixtures in drinking water in response to 

requirements of the Safe Drinking Water Act Amendments of 1996, including 

mixtures of DBPs and of Contaminant Candidate List chemicals (e.g., organotins, 

pesticides, metals, pharmaceuticals). Information and methods are being 

developed to better evaluate the toxic mode of action, the risk posed by drinking 

water mixtures, exposure estimates for mixtures via multiple routes, and the 

relative effectiveness of advanced treatment technologies (USEPA, 2003e, h). 
Page 126 Examination of EPA Risk Assessment Principles and Practices 

b) The Office of Air Quality Planning and Standards used a mixture risk assessment 

approach for multiple route exposures in conducting the National Air Toxics 

Assessment of 33 air pollutants (a subset of 32 air toxics from the Clean Air Act’s 

list of 188 air toxics plus diesel particulate matter) (USEPA, 2001j). 

c) The Office of Air and Radiation issued a proposed rule to evaluate the national 

emission standards for HAPs for stationary combustion turbines, suggesting the 

application of HIs to evaluate potential health risks (USEPA, 2003i). 

d) The Officer of Pesticide Programs conducted a multiple-route assessment of 

organophosphorus pesticide mixtures (USEPA, 2002g), in response to the FQPA 

of 1996. Future assessments may be performed on additional pesticide classes 

and other co-occurring substances for which a common toxic mode of action can 

be identified. 

e) The National Homeland Security Research Center is considering the potential 

toxicological interactions with co-exposure to respiratory toxicants, dust, and 

smoke. 

f) The Office of Solid Waste and Emergency Response evaluates contaminant 

mixtures at Superfund sites (USEPA, 1989a) under CERCLA (see sections 5.7.2 

and 5.7.3 above). Examples of current needs and research areas are the 

assessment of mixtures of metals and guidance on how to group the 

environmental contaminants most commonly found together at sites into common 

mode of action classes for use in developing HIs or assessments based on RPFs. 
Examination of EPA Risk Assessment Principles and Practices Page 127 

6. ECOLOGICAL ASSESSMENT 

6.1 Overview 

The early 1980s saw both the emergence of risk assessment as a regulatory paradigm and 

the first widespread use of ecological impacts to inform regulatory and policy decisions. The use 

of ecological information in decision making has expanded slowly through the 1980s, as 

illustrated by the cancellation of the pesticide diazinon based on impacts to birds (Bascietto, 

1998), the reductions in sulfur dioxide emissions to lessen the adverse impacts of acid deposition 

on lakes and forests and the setting of ozone standards to reduce the damaging effects of 

tropospheric ozone on crops. In the middle to late 1980s, tools and methods for conducting 

ecological risk assessments began to be standardized with the publication of EPA’s Ambient 

Water Quality Criteria (WQC) methodology, the pesticide program’s Standard Evaluation 

Procedures, and Superfund’s Environmental Evaluation Manual (these are discussed in Bascietto 

et al., 1990). More recently, EPA published the Ecological Risk Assessment Guidance for 

Superfund (USEPA, 1997f) and the Guidelines for Ecological Risk Assessment (USEPA, 1998a). 

In 1992, the Agency published the Framework for Ecological Risk Assessment (USEPA, 

1992d; Norton et al., 1992) as the first statement of principles for ecological risk assessment. 

That document and the 1998 first draft Guidelines for Ecological Risk Assessment not only 

describe methods for conducting the more conventional single-species, chemical-based risk 

assessments, but also describe techniques for assessing risks to ecosystems from multiple 

stressors and multiple endpoints. With the publication of these important documents came the 

need to enhance EPA’s ability to do better ecological assessments. These important publications 

recognized the need to advance the science of multiple-scale, multiple-stressor, and multipleendpoint ecological assessments and, more important, called for the consideration of human 

dimensions in the planning and conduct of ecological risk assessments (Bachmann et al., 1998; 

Barton and Sergeant, 1998; Cooper, 1998; Linthurst et al., 2000). 

Following EPA’s paradigm, an ecological risk assessment evaluates the potential for 

adverse ecological effects resulting from human activities. The risk assessment process provides 

a way to develop, organize, and present scientific information so that it is relevant to 

environmental decisions. When conducted for a particular place such as a watershed, the 

ecological risk assessment process can be used to identify vulnerable and valued resources, 

prioritize data collection activity, and link human activities with their potential effects. Risk 

assessments can also provide a focal point for cooperation between local communities and state 

and federal government agencies. Risk assessment results provide a basis for comparing 

different management options, enabling decision makers and the public to make more informed 

decisions about the management of ecological resources. 
Page 128 Examination of EPA Risk Assessment Principles and Practices 

Ecological risk assessments have become a common subject of discourse at EPA. A 

summary of the evolution of this discourse is described by Bascietto et al. (1990), Maki and 

Slimak (1990), Kutz et al. (1992), Norton et al. (1992), Bachmann et al. (1998), and Linthurst et 

al. (2000). Beginning in 1984, the concepts of risk assessment and risk management became a 

common language for justifying regulatory proposals throughout EPA. In 1987 a major Agency 

report stated that the “fundamental mission of EPA is to reduce risks,” and another in 1990 

recommended that EPA “target its environmental protection efforts on opportunities for the 

greatest risk reduction” (USEPA 1987b, 1990). And, as described previously, the rhetoric of risk 

became the Agency’s primary language for justifying its decisions (Russell and Gruber, 1987). 

6.1.1 What Is the EPA Ecological Risk Assessment Approach? 

Ecological risk assessment evolved out of the human health risk assessment process, an 

approach to understanding and protecting humans from the threats — predominately cancer — 

from exposure to chemicals. The National Academy of Sciences’ National Research Council’s 

seminal document on risk assessment (NRC, 1983) defined risk assessment as assessment of the 

“probability that an adverse effect may occur as a result of some human activity.” This report, 

and the articulation by EPA Administrator Ruckelshaus (Ruckelshaus, 1983) that risk was to be 

EPA’s defining operational paradigm, ushered the Agency into a period in which risk assessment 

and risk management became the primary discourse among its staff. Recognizing that its 

definition of risk is human-oriented, the report recommended that the EPA develop a counterpart 

for non-human or ecological endpoints. Thus began a lengthy process within EPA to 

complement the human risk assessment process with procedures for wildlife, ecosystems, and 

endangered species, all of which fall into the category of ecological risk assessment. The EPA 

relied on scientific and technical deliberations among experts in the field to complete the 

Framework for Ecological Risk Assessment (USEPA, 1992d; Norton et al., 1992). The 

Framework defined ecological risk assessment as “a process that evaluates the likelihood that 

adverse ecological effects may occur or are occurring as a result of exposure to one or more 

stressors.” EPA subsequently published case studies (USEPA, 1993b, 1994g) and a companion 

document on issues relating to the elements of an ecological risk assessment (USEPA, 1994h). 

Then (as mentioned above) EPA published an Ecological Risk Assessment Guidance for 

Superfund in 1997 (USEPA, 1997f) and Guidelines for Ecological Risk Assessment in 1998 

(USEPA, 1998a). 

According to EPA (USEPA, 1998a) and others (Suter, 1993; Lackey, 1994), an ecological 

risk assessment is a flexible process for organizing and analyzing data, information, assumptions, 

and uncertainties to evaluate the likelihood of adverse ecological effects. Ecological risk 

assessment provides a critical element for environmental decision making by giving risk 

managers an approach for considering available scientific information along with the other 

factors they need to consider (e.g., social, legal, or economic factors) in selecting a course of 

action. An ecological risk assessment includes an initial planning step, problem formulation, 
Examination of EPA Risk Assessment Principles and Practices Page 129 

analysis of stressors and effects, and risk characterization. Problem formulation, the initial step 

in analysis, involves defining and specifying the issue under consideration. Analysis is separated 

into exposure (stressor) and effects (hazard) analysis. Risk characterization includes both 

quantitative estimation of risk and a qualitative description of risk. The final step is 

communicating the results to a manager and stakeholders. Important properties of the 

assessment include data gathering and review during and after each step and the possibility of 

repeating steps if new data become available. 

Ecological risk assessments at EPA are generally characterized by a particular emphasis 

at the problem formulation phase. The interface among risk assessors, risk managers, and 

interested parties during planning at the beginning of the risk assessment, as well as 

communication of risk at its end, is critical to ensure that the results of the assessment can be 

used to support a management decision. Because of the diverse expertise required, especially in 

complex ecological risk assessments, risk assessors and risk managers frequently work in multidisciplinary teams during problem formulation. This approach has important implications for 

how an ecological risk assessment is formulated, since it requires an up-front discussion on what 

is at risk, what the assessment endpoints are, how they are measured, and what unacceptable risk 

levels are. This discussion and its outcome are shaped by the values, beliefs, and attitudes of the 

risk assessors, risk managers, and stakeholders. Those conducting ecological risk assessments 

have learned that risk managers, risk assessors, and stakeholders bring valuable perspectives to 

the initial planning activities for an assessment. Risk managers charged with protecting the 

environment can identify information they need to develop their decisions, risk assessors can 

ensure that science is effectively used to address ecological concerns, and stakeholders bring a 

sense of realism and purpose. Together, all can evaluate whether a risk assessment can address 

the identified problems. 

6.2 Organism-Level Versus Population-Level Ecological Risk Assessments 

The Agency received comments expressing concern about the use of organism 

(individual) level attributes to assess ecological risk. An example comment follows: 

EPA (1997) ecological risk assessment guidance recommends that potential ecological risks should be 

assesse d at the pop ulation-le vel for a ll but threa tened and e nda ngere d spe cies. [note added here: This 

reference and accompanying citation appear to be incorrect, in that that document does not make such a 

recom mendation.] Although no explicit guidance is pro vided, this is typically accomplished through the use 

of measurement endpoints that are related to population effects (e.g., using Toxicity Reference Values 

based on growth or reproductive effects). However, in many EPA ecological risk assessments, the agency 

has defaulted to assessing effec ts on ind ividual anima ls. This m etric has no sign ificance scientifically and is 

entirely useless as a basis for making risk managem ent decisions. 

The above statement cites text from EPA documents emphasizing the importance of 

protecting populations or communities. The Office of Solid Waste and Emergency Response’s 

guidance for Superfund sites (USEPA, 1999j) states: 
Page 130 Examination of EPA Risk Assessment Principles and Practices 

Eco logica l risk assess men ts incorp orate a wide range of tests and studies to either d irectly estim ate 

com munity e ffects (e.g., b enthic sp ecies d iversity) o r indire ctly pred ict loca l pop ulation-le vel effec ts (e.g., 

toxicity tests on individual species), both of which can contribute to estimating ecological risk. Superfund 

remedial actions generally should not be designed to protect organisms on an individual basis (the exception 

being designated protected status resources, such as listed or candidate threatened and endangered species 

or treaty-protected species that could be exposed to site releases), but to protect local populations and 

communities of biota. Levels that are expected to protect local populations and communities can be 

estimated by extrapolating from effects on individuals and groups of individuals using a lines-of-evidence 

approach. 

As an initial matter, the comment may be more properly considered as a risk management 

decision to be made in the problem formulation stage of an ecological risk assessment. The cited 

text notes the important role that measurements of organism-level effects can play in 

extrapolations to population-level effects. Where possible, toxicity tests are performed using 

measurement endpoints that are related to population effects (e.g., measurements of growth rate 

or reproduction). We agree that it would be desirable to define the full range of ecological 

benefits by considering effects on populations, communities, and ecosystem processes as well as 

organisms. But it is also important to recognize that the use of organism-level attributes such as 

mortality as endpoints for ecological risk assessments is well-supported by policy and precedent 

(USEPA, 1994i, 2003j) and by the courts — e.g., in the diazinon decision (USEPA, 1988c; Ciba 

Geigy v. EPA, 874 F.2d 277 [5th Cir. 1989])1

. In addition, some laws including the Marine 

Mammal Protection Act, the Migratory Bird Treaty Act, The Bald Eagle Protection Act, and the 

Endangered Species Act call for protection of organism properties and even individual 

organisms. Nonetheless, the protection of organism-level attributes is generally interpreted as 

occurring in a population or community context (USEPA, 2003j). That is, increased mortality or 

decreased fecundity or growth of organisms in an assessment population or community is 

assumed to be significant, even with no demonstration that a population- or community-level 

property is affected. 

That assumption is necessitated by the extreme difficulty of predicting effects at higher 

ecological levels. Most population-level attributes, including abundance attributes, are 

determined by the vital rates (births, deaths) of individuals within the population, as well as the 

rates of migration into and out of the population. Linkages between effects on vital rates and 

those on population dynamics, or between effects on the biochemical and physiological processes 

that determine vital rates and effects on populations, can either be established empirically — by 

correlating responses at different levels of biological organization — or by determining causal 

1 On M arch 2 9, 19 88, the Age ncy ca ncelled registratio ns for the pesticid e diaz inon u nless am end ed to 

prohibit use on golf courses and sod farms (USE PA, 198 8c). As noted in the remand decision (Ciba Geigy v. EPA, 

874 F.2d 277 (5th Cir. 19 89) ), the circ uit cou rt rejec ted the industry’s a rgum ent that a r isk is unre ason able o nly if it 

endangers bird populations, stating: “FIFRA gives the Administrator sufficient discretion to determine that recurring 

bird kills, even if they do not significantly reduce bird population, are themselves an unreasonable environmental 

effect.” 
Examination of EPA Risk Assessment Principles and Practices Page 131 

relationships and constructing mechanistic models (see Maltby et al., 2001, for a 

conceptualization and discussion of these linkages). Thus, with some degree of understanding of 

how changes in vital rates manifest into population consequences (i.e., mechanistic 

understanding), effects measured at the organism level can be extrapolated to expected 

population-level responses. Extrapolation of this latter type is usually accomplished using 

models that integrate the effects of stressors on survivorship and fecundity. However, the 

practicality of this approach is cast into question by experience which indicates that so many 

unverifiable assumptions are involved that in real cases the very different outcomes of dueling 

models cannot be resolved (Barnthouse et al., 1984). Also, it is impractical to estimate density 

dependent responses in real populations, particularly when modeling the effects of toxicants. 

The bottom line is that, although methods exist for predicting the effects of chemicals at 

levels of organization higher than the organism, they are still in the development phase and have 

not been shown to be reliable. On the other hand, assessment of ecological risks using measures 

of organism-level effects is justified by experience, policy, and judicial decisions. 

6.3	 Conservatism and Ecological Risk Assessments 

6.3.1	 How Is Conservatism Addressed in Ecological Risk Assessments? 

A number of ecological concerns expressed to the Agency revolve around the concept of 

being overly conservative. In early tiers of a risk assessment (e.g., for screening), a high degree 

of conservatism is sought. Definitive assessments strive to be as realistic as possible, replacing 

conservative assumptions with best estimates of exposures and effects and associated 

uncertainties. Three examples in which conservatism is addressed in ecological risk assessments 

are discussed in the following sections. 

6.3.2	 How Are Toxicity Reference Values Developed in Ecological Risk 

Assessments? 

A number of comments have been received relating to the development of toxicity 

reference values (TRVs) for use in ecological assessments. One of these notes that EPA 

guidance recommends use of a “weight-of-evidence” method that does not generate TRVs 

appropriate for individual sites. The guidance that this comment refers to is the Ecological Soil 

Screening Levels (Eco-SSLs) document (USEPA, 2003k). The intent of this document is 

precisely to develop generic TRVs that are not site-specific: the TRVs are meant to be 

conservative and to be used for screening purposes. 

In the aquatic arena, the use of conservative threshold concentrations for ecological 

protection is not new. EPA Water Quality Criteria (WQC), designed to protect aquatic 

organisms and their uses — specifically, 95% of the taxa from an appropriate variety of 
Page 132 Examination of EPA Risk Assessment Principles and Practices 

taxonomic and functional groups (USEPA, 1985) — have been employed in a regulatory context 

for many years. The degree to which Eco-SSLs (derived from generic TRVs) may be more 

conservative than chronic WQC can not be assessed, because of differences in their derivation. 

Another comment was received relating to inconsistency in use of UFs in developing 

avian TRVs for dioxin. At issue are the conditions under which a subchronic-to-chronic UF 

should be used. Use of an UF where unwarranted would result in an overly conservative TRV. 

Based on guidance in Sample et al. (1996), avian studies where exposure duration is 10 weeks or 

less are generally considered to be subchronic. Sample et al., however, considered 10-week 

exposure of ring-necked pheasant hens to TCDD in a study by Nosek et al. (1992a) to represent 

chronic conditions, in part because exposure occurred through a critical life-stage (reproduction). 

It must be noted, though, that TCDD’s persistence and bioaccumulative properties make it an 

exception to the rule that exposure during a critical life-stage is sufficient for consideration as 

chronic exposure. Another study by Nosek et al. (1992b) found a half-life for whole-body 

elimination of TCDD in non egg-laying pheasant hens of approximately 1 year. Based on this 

information, only 13% of steady-state accumulation would be achieved from a 10-week 

exposure. A truly chronic exposure could presumably have had nearly an order of magnitude 

lower concentration in the food and still elicited the same tissue levels and effects (USEPA, 

1993c). Thus, the 10-week exposure in the Nosek et al. study (1992a) can be considered to be 

subchronic, and a UF of 10 used to derive a TRV. In summary, all available information is taken 

into account when deciding on the use of an UF. 

6.3.3	 How Do Screening-Level and Definitive (Baseline) Assessments Differ? 

The Agency has been criticized for using screening-level ecological risk assessments 

rather than definitive “baseline” assessments in making remedial decisions. The former are 

designed to be conservative, and are generally meant to be followed by more detailed assessment. 

A screening-level ecological risk assessment provides a comparison of abiotic media 

concentrations to ecotoxicological benchmarks, and does not include extensive site-specific 

information. In a definitive assessment, exposure concentrations are derived for wildlife 

receptors as well as for abiotic media. Definitive assessments require analysis of exposureresponse relationships for the chemicals of concern. The use of site-specific information to 

derive appropriate exposure concentrations is an important element of a definitive assessment 

and is discussed below. 

6.3.4	 How Is Site-Specific Information Used To Derive Exposure Concentrations 

in Ecological Risk Assessments? 

The Agency has been criticized for ignoring site-specific habitat information in a number 

of ecological risk assessments for Superfund sites. The degree to which spatial and temporal use 

of habitats by receptor species should be considered in the conceptual site model is an important 

question. Distinct management areas, known as operable units or OUs, are often evaluated in the 
Examination of EPA Risk Assessment Principles and Practices Page 133 

RI/FS process. It is important to have an understanding of the degree to which various habitats 

within an OU are, or will be, used by receptor species. For example, if we know that fish move 

throughout the area of an OU, exposure concentrations should be derived using data from the 

entire OU. Exposure estimates are generated using point estimates of exposure parameters (e.g., 

upper 95 % confidence limits of the mean chemical concentrations in media). More realistic 

estimates of exposure may be generated using distributions of exposure parameters and 

uncertainty analysis (e.g., Monte Carlo simulation; see USEPA, 1997b, and the references in 

chapter 7 of Suter et al., 2000). In the case of a large river system with tributaries that may be 

“relatively” clean compared to the main channel, the decision as to whether or not to include 

tributaries in the assessment typically depends on how contaminated they are and how they are 

used by the receptor species. 

6.4	 Water Quality Criteria 

6.4.1	 What Is the Agency Standard Methodology for Deriving Aquatic Life Water 

Quality Criteria? 

Criteria are derived according on the principles set forth in the 1985 Guidelines for 

Deriving Numerical National Water Quality Criteria for the Protection of Aquatic Organisms 

and Their Uses (USEPA, 1985). These guidelines are ordinarily applied as follows: 

a)	 Acute toxicity test data need to be available for species from a minimum of eight 

diverse taxonomic groups. The diversity of tested species is intended to ensure 

protection of various components of an aquatic ecosystem. 

b)	 The Final Acute Value (FAV) is derived by extrapolation or interpolation to a 

hypothetical genus more sensitive than 95% of all tested genera. The FAV, which 

typically represents an LC50 or EC50, is divided by two in order to obtain an acute 

criterion protective of nearly all individuals in such a genus. 

c)	 Chronic toxicity test data (representing longer-term survival, growth, or 

reproduction end points) need to be available for at least three taxa to be 

calculated directly. When such data are not available, which is often the case, the 

chronic criterion is set by determining an appropriate acute-chronic ratio (the ratio 

of acutely toxic concentrations to chronically toxic concentrations) and applying 

that ratio to the acute value of the hypothetical genus more sensitive than 95% of 

all tested genera. 

d)	 When necessary, the acute and/or chronic criterion may be lowered to protect 

recreationally or commercially important species. 
Page 134 Examination of EPA Risk Assessment Principles and Practices 

6.4.2	 What Is the GLI Tier II Methodology? 

The Tier II methodology for aquatic life is contained in the final Water Quality Guidance 

for the Great Lakes System, or Great Lakes Initiative (GLI). The methodology was developed for 

use in deriving translators for narrative criteria to protect aquatic life in waters of the Great Lakes 

system when insufficient data are available to derive criteria. Tier II values may be based on 

toxicity data from as little as a single taxonomic family, provided the data are acceptable. The 

values are generally more stringent than criteria to reflect greater uncertainty in the absence of 

additional toxicity data. As more data become available, the Tier II values tend to become less 

conservative and more closely approximate the criteria. The general procedure is: 

a) The lowest Genus Mean Acute Value in the database is divided by the Secondary 

Acute Factor (an empirically derived adjustment factor that varies according to the 

number of satisfied minimum data requirements listed in the criteria 

methodology) to derive the Secondary Acute Value (SAV). 

b) A Secondary Chronic Value (SCV) is calculated from either the FAV (from the 

criteria) or the SAV divided by either the Secondary Acute to Chronic Ratio or 

Final Acute to Chronic Ratio. 

c) The Secondary Maximum Concentration is equal to half of the SAV, and the 

Secondary Continuous Concentration is equal to the SCV or the Final Plant 

Value, if available — whichever is lower. 

d) When necessary, the acute and chronic values can be adjusted to protect 

recreationally or commercially important species. 

The final rule allows and encourages dischargers, states, and authorized tribes to develop 

additional data that could reduce the Tier II adjustment factors or allow development of WQC 

(Tier I criteria) using the methodology described above. Additionally, in situations when 

dischargers generate data to derive Tier I criteria or Tier II values, the final rule specifies that 

permit authorities may grant a reasonable period of time for data generation before the criteria or 

values become effective for the discharge permit. 

6.4.3	 Was the GLI Tier II Methodology Made Available for Public Comment or 

Vetted Through Other Public Processes? 

The Tier II methodology was first developed by a Steering Committee, consisting of 

directors of water programs from EPA’s national and regional offices and the Great Lakes states’ 

environmental agencies (as co-regulators of Clean Water Act water quality programs), and a 

Technical Work Group consisting of technical staff from the Great Lakes states’ environmental 

agencies, EPA, the U.S. Fish and Wildlife Service, and the U.S. National Park Service. All of 
Examination of EPA Risk Assessment Principles and Practices Page 135 

the work was reviewed and commented upon by a Public Participation Group consisting of 

representatives from environmental groups, municipalities, industry, and academia who observed 

the deliberations of the other two groups, advised them of the public’s concerns, and kept their 

various constituencies informed. 

EPA’s SAB reviewed the draft Tier II methodology in 1992. (See its final report, 

EPA-SAB-EPEC/DWC-93-005, completed December 16, 1992.) The SAB commended EPA for 

the interactions among the states, EPA, the private sector, and the scientific community during 

the development of this proposed guidance and provided substantial comments on many 

elements of the submitted draft documents. The Board provided specific comments on the Tier II 

methodology. EPA responded to each of these comments in proposing and finalizing the 

methodology. The preamble to the proposed rule (see below) describes these comments and 

EPA’s response to each. 

The Tier II methodology was part of a 1995 rulemaking (USEPA, 1995c, d) and as such 

received full public comment, response, and revision based on the comments, in accordance with 

the Administrative Procedure Act. Federal Register notices requesting public comment on the 

guidance were published on April 16, 1993, August 9, 1993, September 13, 1993, and August 30, 

1994 (e.g., see USEPA,1993d). Over 26,500 pages of comments, data, and information from 

over 6,000 respondents were received in response to these Federal Register notices and meetings 

held with the public. 

6.5	 Uncertainty and Pesticide Ecological Assessment 

6.5.1	 How Does EPA Address Uncertainty Analysis in Pesticide Ecological 

Assessments? 

As a result of a dialogue with internal and external experts in the field of ecological risk 

assessment and probabilistic methods, the Environmental Fate and Effects Division of the Office 

of Pesticide Programs is currently developing methodology to routinely conduct probabilistic risk 

assessments (http://www.epa.gov/oppefed1/ecorisk/index.htm). The models were initially 

presented to the FIFRA Scientific Advisory Panel in March 2001 

(http://www.epa.gov/scipoly/sap/2001/index.htm#march). The Panel recommended a number of 

improvements prior to implementation; these improvements will be discussed with the Panel in a 

SAP meeting originally scheduled for February 2004 followed by implementation in fiscal year 

2004. The models were used in an assessment of the risks of the use of carbofuran flowable 

which is currently scheduled for completion in fiscal year 2004. In the interim, pesticide 

registrants have submitted probabilistic exposure assessments for chlorpyrifos, chlorfenapyr, 

diazinon, and atrazine. The chlorfenapyr assessment was reviewed by the FIFRA SAP, which 

concluded that the registrants’ conclusion of negligible risk is not supported for any geographic 

scale. The scaling up of field exposures to the Cotton Belt scale was done incorrectly. The 

probabilistic exposure assessment had several other flaws that require remediation (see 
Page 136 Examination of EPA Risk Assessment Principles and Practices 

http://www.epa.gov/scipoly/sap/1999/september/finalrpt.pdf). The Environmental Fate and 

Effects Division conducted a modified probabilistic assessment for endosulfan in which a 

distribution of effects was overlaid on a distribution of exposure to generate a joint probability 

curve. 

Registrants have indicated that probabilistic risk assessments are planned for a 

rodenticide, while the Division may conduct one for a registered chemical going through the reregistration process in early fiscal 2004. 

6.5.2 How Does OPP Address Rodenticide Use Patterns? 

EPA issued the document Potential Risks of Nine Rodenticides to Birds and Nontarget 

Mammals: A Comparative Approach in January 2003 for public comment (USEPA, 2003l). The 

Agency is currently reviewing the comments submitted to OPP’s docket. Similar comments 

were submitted in response to OMB’s February 3, 2003, request for public comments on its Draft 

2003 Report to Congress on the Costs and Benefits of Federal Regulations (USOMB, 2003a). 

The document presented the Agency’s preliminary assessment of potential risks to birds 

and non-target mammals from nine rodenticides, predominantly used to control commensal rats 

and mice in and around buildings and transport vehicles, and in sewers. Some also have products 

registered for other outdoor uses against other rodent and small mammalian pests. The 

assessment focused on the potential primary and secondary risks to birds and non-target 

mammals posed by applications of these nine rodenticides (11 baits) to control rats and mice in 

and around buildings (commensal use) and in field and other outdoor settings to control various 

rodent and other small mammalian pests. Typical use information used to estimate non-target 

organism exposure, such as amount of rodenticide active ingredient or formulated product 

applied per unit area, was not available. Thus, exposure estimates are largely based on the 

amount of active ingredient available per kilogram of the grain bait formulation. In preliminary 

risk assessments, it is assumed that birds and non-target mammals are likely to be exposed to the 

pesticide without attempting to establish a quantitative measure of likelihood. The existence of 

incident data along with liver residues provides important support for the assumption that birds 

and non-target mammals are exposed and adversely affected by applications of rodenticide baits. 

The fact that numerous species, including predators and scavengers, have been found exposed to 

these baits indicates that both primary and secondary exposures are occurring. 

The risk conclusions were based both on the weight of evidence of the available data and 

comparative analysis modeling. Each rodenticide is ranked or categorized and compared to the 

other rodenticides according to the following criteria: 

a) Overall potential risk. 

b) Potential primary risk to birds. 
Examination of EPA Risk Assessment Principles and Practices Page 137 

c)	 Potential primary risk to non-target mammals. 

d)	 Potential secondary risk to avian predators and scavengers. 

e)	 Potential secondary risk to mammalian predators and scavengers. 

A comment has been received that suggests the rodenticide comparative assessment fails 

to evaluate the likelihood of ecological exposure for the many different use patterns in the 

preliminary comparative “ecological risk” assessment of rodenticides. 

A number of factors contributed uncertainty to the assessment. These were discussed in 

the document released to the public. Those that appeared to contributed the greatest uncertainty 

were: 

a)	 Missing data on acute, chronic, and secondary effects, as well as retention of some 

active ingredients in the liver and blood. 

b)	 The variable quality and quantity of data on metabolism and retention times in 

rodents and non-target species. 

c)	 Specific use information by formulation, including typical amounts applied, 

distances applied from buildings, amounts used in rural versus urban areas, and so 

forth. 

d)	 Information on the number and species of birds and non-target mammals likely to 

find and consume bait in the various use areas. 

e)	 Methods to determine liver concentration(s) that would corroborate death from 

anticoagulant exposure, or even if such a cause-effect relationship is appropriate, 

e.g., the “threshold of toxicity” concentration in liver tissue. 

f)	 Not accounting for the impacts of sub-lethal effects on non-target mortality, e.g., 

clotting abnormalities, hemorrhaging, and stress factors (including environmental 

stressors such as adverse weather conditions, food shortages, and predation). 

g)	 Comparing rodenticides with different modes of action, i.e., vitamin K antagonists 

that disrupt normal blood-clotting (anticoagulants), a diphenylamine that is a 

neurotoxicant, an inorganic compound that kills by liberation of phosgene gas, 

and a sterol that kill by inducing hypercalcemia. 
Page 138 Examination of EPA Risk Assessment Principles and Practices 

A comparative analysis model was used to rank and compare potential primary and 

secondary risks. The underlying methodology is a simple multi-attribute rating technique, or 

SMART (Goodwin and Wright, 1998). SMART was adapted for comparing potential risks 

among rodenticides based on a number of measure-of-effect values for primary and secondary 

risk to birds and mammals. It is similar to the technique used in the Agency’s Comparative 

Analysis of Acute Avian Risk From Granular Pesticides (USEPA, 1992e) and A Comparative 

Analysis of Ecological Risks From Pesticides and Their Uses: Background, Methodology, and 

Case Study (USEPA, 1998e); both were reviewed by a FIFRA Scientific Review Panel. 

Concerning the latter analysis, the Panel noted the many scientific uncertainties in the method, 

yet agreed that it was a useful screening tool that provides a rough estimate of relative risk. The 

Panel made a number of helpful suggestions to improve the utility of the method, most of which 

are included in this comparative assessment. In this analysis, a risk quotient (RQ), calculated as 

the ratio of toxicant potentially ingested to the inherent toxicity of the rodenticide, was used to 

compare potential primary risks to birds and non-target mammals. RQs are compared among 

rodenticide baits based on the amount of bait and number of bait pellets that birds or non-target 

mammals of various sizes would need to eat to ingest an acute oral (LD50) dose. Dietary data 

(LC50) also are available for birds (but not for mammals), and RQs based on bait concentration 

and avian dietary toxicity are compared among the rodenticides. As noted by the Ecological 

Committee on FIFRA Risk Assessment Methods (ECOFRAM, 1999), RQs do not quantify risk 

but are useful for comparisons among alternative compounds. EPA’s Guidelines for Ecological 

Risk Assessment (USEPA, 1998a) also notes that quotients provide an efficient, inexpensive 

means of identifying high- or low-risk situations that can allow risk management decisions to be 

made without the need for further information. 

A major uncertainty has been how and where these pesticides are used. In estimating 

pesticide exposure, risk assessors rely upon the label instructions, since the label prescribes the 

level and manner of use. Use other than that specific on the label is illegal. The commensal use 

is common to all nine rodenticides. The terminology “in and around buildings” appears on 

product labels registered for commensal use. This statement does not limit bait placements to 

any specified distance from buildings, and in many non-urban areas bait applications might pose 

an exposure scenario comparable to some field uses. Only two of the nine rodenticide labels 

limit the “in and around buildings” use to urban areas; applications in non-urban areas must be 

indoors. Because of this label language, exposure to non-targets has been difficult to estimate: 

the phrase “in and around buildings” — i.e., how far from a building bait may be applied — is 

subject to interpretation. Further, many of these products are sold to the general public and 

available in outlets ranging from grocery to hardware stores. Homeowners may or may not 

follow the label instructions. How and where the public and pest control operators use these 

products is largely unknown. Currently, general marketing data are available. At best, these data 

report sales on a regional basis, and thus may be used to estimate volume of use. California does 

require that reports be maintained of pesticide use, but homeowners are exempt. 
Examination of EPA Risk Assessment Principles and Practices Page 139 

For agricultural uses, the Agency obtains information on how the pesticide is typically 

used: application rates, number of applications, areas in the United States where it is used, etc. 

The registrants, the Rodenticide Registrants Task Force, have volunteered to provide usage 

information for the rodenticide of most concern which will allow the Agency to further refine its 

risk assessment. 

6.6 Summary and Conclusions 

This chapter examines ecological risk assessment approaches in a few areas within the 

Superfund, water, and pesticides programs in order to address issues raised by comments 

received by EPA. Within the Superfund program, we discuss EPA’s approach regarding 

organism- versus population-level ecological risk assessments and provide three examples in 

which conservatism is addressed. Within the water program, we discuss the standard 

methodology for deriving aquatic life WQC and the Great Lakes Initiative Tier II Methodology. 

In the pesticides program we discuss how EPA addresses uncertainty analysis in pesticide 

ecological assessments and rodenticide use patterns. To paraphrase a statement made earlier in 

this document, we attempt to use all information available in an objective, realistic, scientifically 

balanced way to make decisions. 

Ecological risk assessment “evaluates the likelihood that adverse ecological effects may 

occur or are occurring as a result of exposure to one or more stressors.” It is a flexible process 

for organizing and analyzing data, information, assumptions, and uncertainties to evaluate the 

likelihood of adverse ecological effects. Ecological risk assessment provides a critical element 

for environmental decision making by giving risk managers an approach for considering 

available scientific information along with the other factors they need to consider (e.g., social, 

legal, political, or economic) in selecting a course of action. 

Ecological risk assessment includes three primary phases: problem formulation, analysis, 

and risk characterization. In problem formulation, risk assessors evaluate goals and select 

assessment endpoints, prepare the conceptual model, and develop an analysis plan. During the 

analysis phase, assessors evaluate exposure to stressors and the relationship between stressor 

levels and ecological effects. In the third phase, risk characterization, assessors estimate risk 

through integration of exposure and stressor-response profiles, describe risk by discussing lines 

of evidence and determining ecological adversity, and prepare a report. 

The interface among risk assessors, risk managers, and interested parties during planning 

at the beginning of the risk assessment, and communication of risk at its end, is critical to ensure 

that the results of the assessment can be used to support a management decision. Because of the 

diverse expertise required (especially in complex ecological risk assessments), risk assessors and 

risk managers frequently work in multidisciplinary teams. Both risk managers and risk assessors 

bring valuable perspectives to the initial planning activities for an ecological risk assessment. 

Risk managers charged with protecting the environment can identify information they need to 
Page 140 Examination of EPA Risk Assessment Principles and Practices 

develop their decisions, risk assessors can ensure that science is effectively used to address 

ecological concerns, and together they can evaluate whether a risk assessment can address 

identified problems. However, this planning process is distinct from the scientific conduct of an 

ecological risk assessment. This distinction helps ensure that political and social issues, while 

helping to define the objectives for the risk assessment, do not introduce undue bias. 

Problem formulation, which follows these planning discussions, provides a foundation 

upon which the entire risk assessment depends. Successful completion of problem formulation 

depends on the quality of three products: assessment endpoints, conceptual models, and an 

analysis plan. Since problem formulation is an interactive, nonlinear process, substantial 

reevaluation is expected to occur during the development of all problem formulation products. 

The analysis phase includes two principal activities: characterization of exposure and 

characterization of ecological effects. The process is flexible, and interaction between the two 

evaluations is essential. Both activities evaluate available data for scientific credibility and 

relevance to assessment endpoints and the conceptual model. Exposure characterization 

describes sources of stressors, their distribution in the environment, and their contact or cooccurrence with ecological receptors. Ecological effects characterization evaluates stressorresponse relationships or evidence that exposure to stressors causes an observed response. The 

bulk of quantitative uncertainty analysis is performed in the analysis phase, although uncertainty 

is an important consideration throughout the entire risk assessment. The analysis phase products 

are summary profiles that describe exposure and the stressor-response relationships. During risk 

characterization, the final phase of an ecological risk assessment, risk assessors estimate 

ecological risks, indicate the overall degree of confidence in the risk estimates, cite evidence 

supporting the risk estimates, and interpret the adversity of ecological effects. To ensure mutual 

understanding between risk assessors and managers, a good risk characterization will express 

results clearly, articulate major assumptions and uncertainties, identify reasonable alternative 

interpretations, and separate scientific conclusions from policy judgments. 

The Agency’s Ecological Risk Assessment Guidelines are just that — guidelines, meant 

not to be overly prescriptive. Each program office administers different statutes that in many 

cases are silent on assessment endpoints or that stipulate endpoints that can be difficult to 

quantify. As such, programs have the responsibility to develop assessment procedures that are 

consistent with their enabling legislation. For the most part, the various programs’ practice of 

ecological risk assessment is consistent across the Agency, but there are enough differences to 

raise questions of consistency, as illustrated by the comments this document seeks to address. 

EPA’s goal is to apply ecological risk assessment methods in as consistent and transparent a 

process as possible, enabling our stakeholders to engage in a fully open debate on the risks and 

the management of those risks for the overall protection of the environment. 
Examination of EPA Risk Assessment Principles and Practices Page 141 

7. SUMMARY AND RECOMMENDATIONS 

7.1 Summary 

EPA’s mission and statutes require that EPA protect human health and the environment, 

and ensure that EPA’s risk management decisions provide some “margin of safety” when risk 

assessment answers are uncertain or variable. The risk estimates provided in risk assessments 

and ultimately addressed in EPA decisions are driven by the uncertainty and variability inherent 

in practically all the information and methodologies EPA uses. This uncertainty stems to a large 

extent from the complex nature of the environment and the processes that interact with the 

stressor(s) to cause risk. It is clear that risk assessors best understand the basic elements of a risk 

assessment, in particular how to analyze the uncertainty and variability found in individual risk 

assessments. Risk managers, by taking into consideration the characterization of all the elements 

of a risk assessment — including the uncertainty and variability, provided by the risk assessors 

— should determine the appropriate degree of protection to ensure that risk is not underestimated 

and conversely, is not appreciably overestimated. 

Comments on EPA risk assessment principles and practices have been received over the 

years, more recently in response to the OMB Federal Register notice asking for such comments. 

An overarching position in the comments suggests that EPA inappropriately mixes policy and 

science: that, in generating risk assessments, risk assessors “decide” the degree of protection in a 

way that is not transparently characterized. Further, in more than a few instances, it is suggested 

that EPA’s current risk assessment practice compounds risk estimates to a protective position 

that is beyond reasonable. Also, in many instances, risk managers apparently are not aware of 

the degree of protection inherent in each risk assessment. Some other comments suggest that 

EPA risk estimates tend to underestimate risk, since a majority of the assessments tend to be 

single-stressor assessments and do not adequately take into account multiple stressors, multiple 

exposures, and susceptible populations. 

EPA risk assessment practices are meant to produce credible, science-based risk 

assessments that provide reasonable risk estimates. The assessments are designed to ensure that 

risk is not underestimated or grossly overestimated. EPA views a critical analysis of the 

available data specific to the relevant chemical (stressor) and/or site as the starting point for a 

risk assessment. The derivation of risk estimates improves continually with the addition of 

newer techniques and relevant data. In the absence of relevant data, EPA uses assumptions in 

developing estimates of risk. The assumptions (defaults and extrapolations) used in the 

assessments are necessary components of the assessment, as they help bridge the data gaps 

encountered during the assessment. These assumptions are also focused on providing a 

reasonable default position and usually have scientific data and scientific support. They are also 

publicly vetted and peer reviewed. For the conduct of risk assessment, EPA uses a triage 

approach (from screening to “in-depth” assessments) to determining how much time and 
Page 142 Examination of EPA Risk Assessment Principles and Practices 

resources need to be expended. This is a practical approach for dealing with the generation of 

different “depths” of assessment. 

Another major theme in the comments regarding EPA risk assessment principles and 

practices suggests that transparency and clarity are still absent in characterizing the elements of 

the risk assessment, particularly the uncertainty and variability as well as the choices made in the 

assessment. This may be an aspect of EPA’s practices that needs strengthening. It should be 

clear that the development of risk assessments for risk management purposes is a mix of science 

data, science policy, judgments, guidelines, and best professional judgement. Transparency 

about all these matters is crucial so that nothing appears “hidden” or “buried” in an assessment 

— so that nothing keeps one from understanding the impact of the elements that go into 

estimating and characterizing a risk. 

7.2	 Several Current EPA Activities That Will Enhance EPA Risk Assessment Principles 

and Practices 

EPA works on many fronts to review our practices and methodologies as part of an 

ongoing evaluation of risk assessment principles and practices. Several current examples 

demonstrate our commitment to enhancing EPA risk assessment principles and practices: 

a)	 EPA is currently finalizing its proposed Guidelines for Cancer Risk Assessment. 

Included in this effort is the generation of a supplemental guidance focused on 

cancer risk to children, particularly from mutagenic carcinogens. This is a major 

effort that explores many newer issues, such as examining data before invoking 

defaults, extensive use of mode of action to inform the assessment, more clearly 

defining default positions that can be used when data are unavailable, and making 

the cancer risk assessment practice more transparent and clear. 

b)	 EPA is engaged in a long-term examination to “harmonize” cancer and noncancer health assessments into a unified approach more geared toward examining 

how health endpoints, via a mode of action, present linear or nonlinear doseresponse curves that can be used for quantitation. Other major aspects of this 

effort are the use of chemical-defined uncertainty factors, guidance for selecting a 

point of departure from observed data to extrapolate to lower doses of expected 

exposures, and clarification of dosing and scaling factors to use in health 

assessments. 

c)	 The use of environmental models to support risk assessments is constantly 

increasing. EPA promotes consistency and consensus on mathematical modeling 

issues including modeling guidance, development, and application, and enhances 

both internal and external communications on modeling activities through 

organizations such as the EPA Council on Regulatory Environmental Modeling. 
Examination of EPA Risk Assessment Principles and Practices Page 143 

d)	 The better the data and information are, the better the risk assessment will be in 

terms with being more reasonable with more certainty. To that end, EPA initiated 

an effort to examine the usefulness of the toxicity tests needed, and required under 

several statutes, to support EPA assessments and decisions. As the field of 

toxicology evolves, particularly recently with the Human Genome Project and the 

burgeoning area of genomics and related technologies, it is appropriate to examine 

what will be the best tests we can use to generate data useful for risk assessments. 

EPA is working with the NAS to examine these toxicity testing needs as projected 

into the next two decades. 

e)	 EPA is undertaking a major effort to upgrade and improve its process for creating 

assessments to enter onto IRIS. As a result of an increase in IRIS staff and 

funding, the number of new and updated assessments completed each year will be 

increased. A more open external peer review component will provide more 

transparency and interaction for external stakeholders. 

f)	 EPA focuses much attention on its research program to help reduce major 

uncertainties in risk assessment by understanding and elucidating the fundamental 

determinants of exposure and dose and the basic biological changes that follow 

exposure to stressors leading to a toxic response. For example, in EPA’s recent 

Human Health Research Strategy (USEPA, 2003m), EPA will focus on ways to 

improve the information supporting its risk assessments. These include 

approaches to harmonizing the use of mechanistic data in human health 

assessments, predicting the effects of aggregate and cumulative exposure, and 

protecting susceptible populations such as children, older adults, and those with 

preexisting disease or genetic predispositions for different responsiveness to 

environmental stressors. Many of these areas are discussed in this document. 

7.3	 Recommendations for Further Improvements to Risk Assessment Principles and 

Practices 

Based on the evaluation in this document, the staff provides several recommendations for 

EPA to consider as part of an agenda for further improvements to risk assessment principles and 

practices. 

a)	 Confidence in our risk assessments is critical. This confidence is increased by the 

use of scientifically sound, chemical- and site-specific information. In the 

absence of such data, EPA should continue to use default assumptions to fill data 

gaps. 
Page 144 Examination of EPA Risk Assessment Principles and Practices 

1)	 We need to continue to encourage the development of the specific data 

necessary to more accurately assess potential risks, particularly with those 

responsible for generating appropriate data (e.g., pesticide registrants). In 

the absence of strong data requirements, EPA should continue to work 

with industry and other partners to implement voluntary programs, such as 

the High Production Volume Challenge, to generate such data, thus 

lessening our reliance on defaults in developing risk assessments. 

2)	 Since defaults assumptions are used in risk assessments to address 

uncertainty when chemical- and/or site-specific data are not available or 

useful, we need to ensure that the defaults themselves are supported by the 

best available data. For example, the draft supplemental children’s 

guidance accompanying the EPA draft cancer guidelines suggests a default 

position for children exposed to mutagenic carcinogens that is based on 

data, albeit with a small sample number, that supports the default 

assumption (USEPA, 2003d). We should continually look for 

opportunities to increase our certainty and confidence in the defaults and 

extrapolations we use. As more data become available, we develop and/or 

refine the default assumptions and extrapolations to use; for example, 

defaults might be developed for different life-stages or for chemicals with 

a specific mode of action. EPA should encourage and support research 

needed to develop and refine assumptions for use in risk assessments. 

Further, to help increase our confidence in risk assessments, EPA might 

consider the use of sensitivity analysis to examine what data points and/or 

defaults involved in the assessment are the most critical. 

b)	 Transparency in our risk assessment practice and our risk management process is 

an issue that needs to be addressed. Initiatives for EPA to improve our 

transparency in risk assessments and risk management include: 

1)	 Continued concerted and conscious use of planning and scoping with risk 

assessors and risk managers before a risk assessment is started. The intent 

is not to point to a specific decision choice, but to outline what needs to be 

addressed and how to do so. Just as important is to be clear about what 

will not be addressed (i.e., the scope of the assessment effort needs to be 

defined). 

2)	 Continued use of the triage approach to decide how much time and 

resources are necessary for a risk assessment. However, we need to be 

more clear about why the risk assessment is being performed. This would 

help reduce potential confusion about why EPA decides upon particular 

risk estimates in one decision and others in another decision (e.g., a 
Examination of EPA Risk Assessment Principles and Practices Page 145 

screening assessment versus a detailed assessment; different circumstances 

at different sites). 

3)	 We need to focus on better communication of the data and assumptions 

and choices used in our risk assessments. We already have the guidance in 

place in the form of our Risk Assessment Guidelines and Agency Risk 

Characterization Policy and Handbook. Also, we have Planning and 

Scoping Guidance as the first part of our Cumulative Risk initiative, and 

this is incorporated into the new Cumulative Risk Framework. Close 

attention to our guidances will ensure transparency and clarity in our risk 

assessments. This will also help ensure that risk assessments will 

appropriately support and inform decision making and risk management. 

4)	 Regarding transparency in the risk management process: since we do not 

have much guidance here, we should encourage work on a decision 

making framework. This framework would help everyone, from those 

inside EPA to all interested parties outside EPA, understand the process 

risk managers (decision makers) go through to arrive at a decision. It 

should not be a checklist of steps, but a set of actions (e.g., peer review) 

and factors (e.g., cost/benefit analysis) risk managers should consider 

before making a decision. While it is certainly feasible to produce such a 

framework explaining our methods of procedure in general terms, it 

cannot explain relevant “weights” given in risk management decisions to 

the various factors (costs, benefits, treatment options, etc.) as each 

decision will be made on a case-by-case basis. So most importantly, this 

transparency will clarify how a risk assessment is incorporated into a 

decision and how it informs the decision. Places to look for information to 

construct such a framework include the NAS reports and the Commission 

on Risk Assessment and Risk Management. 

c)	 A major method to help address uncertainty in risk assessment is probabilistic 

analysis. We can use that technique for exposure assessments, but even for these 

assessments, we should encourage greater use and reliance on it when appropriate. 

Further, we need to explore the feasibility of using probabilistic analyses in any 

phase of the risk assessment. For example, we need to encourage the 

development of Monte Carlo and other probabilistic analyses in the dose-response 

assessment. If probabilistic analyses are identified for particular uses and it is 

generally agreed they are appropriate, we should start incorporating the relevant 

analyses where the risk assessor feels it is appropriate to do so. 

1)	 It should be noted that the seemingly slow implementation of probabilistic 

analysis is not strictly an internal EPA issue. Much environmental risk 
Page 146 Examination of EPA Risk Assessment Principles and Practices 

assessment is performed by affected parties rather than EPA itself, and 

outside parties’ pace in implementing and/or accepting 

probabilistic/quantitative analysis of uncertainty is seemingly slow as well. 

Notwithstanding, we recommend that EPA examine a few recent or 

ongoing risk assessments to determine what the obstacles are to use of 

probabilistic analysis to describe uncertainty and how to overcome those 

obstacles more frequently in the future. A major area that needs work is 

the development of probability distributions for many endpoints in the 

toxicity part of risk assessment, as well as putting the toxicity and 

exposure components together in an overall probabilistic risk assessment. 

Preliminary work in this area has been published, with some key papers 

concerning the development of probability distributions for uncertainty 

factors (Dourson et al., 1992; Baird et al., 1996; Swartout et al., 1998). 

d)	 The staff feels greater harmonization with our partners — states, other national 

governments (as in the case of OPP’s interactions with Canada), OECD — as well 

as greater cross-program interaction inside the Agency to ensure consistency and 

scientific consensus for practices and methodologies (e.g., probabilistic analyses) 

will help avoid a duplication of effort or differences in approaches and 

interpretations. 

e)	 EPA has historically assessed risks based on individual chemicals/stressors and 

often focused on one source, pathway, or adverse effect. However, people do not 

live in a “single chemical” world: in reality, they are exposed to multiple 

chemicals/stressors from a variety of sources. Because people are exposed to 

multiple stressors, this suggests that the total risk they may experience in their 

daily lives may be underestimated by assessments of single stressors. Tools are 

needed to understand combined risks. While the Agency has moved forward with 

emphasis on evaluating cumulative risks (e.g., planning and scoping activities for 

cumulative risk, the Cumulative Risk Framework, OPP examination of aggregate 

risk), EPA needs to continue moving aggressively to flesh out approaches for 

cumulative risk and to produce the most scientifically rigorous and realistic 

evaluation of cumulative risk that the state-of-the-science can accommodate. 

f)	 Overarching guidance is needed to improve the ease and consistency of risk 

assessment for susceptible populations and life-stages. Such guidance should 

provide more detail, methods, and advice for considering the issues regarding 

susceptibility. As a part of such efforts, EPA needs to continue developing a 

strategy and methodologies to address the cumulative risks of unique or 

disproportionate exposures. 
Examination of EPA Risk Assessment Principles and Practices Page 147 

g) Peer review is fundamental to ensuring the quality and integrity of all 

assessments. We need to continue with constant vigilance to peer review all our 

major scientific and technical work products. 

h) Many practices EPA is criticized for are practices we no longer follow (e.g., 

compound uncertainty factors up to 10,000). However, there are older risk 

assessments supporting standing decisions that may need reexamination. This 

would need to be done in light of current resources and priorities, e.g., looking at 

older IRIS values for chemicals that are still EPA priorities. 
Page 148 Examination of EPA Risk Assessment Principles and Practices 

LIST OF USEFUL ABBREVIATIONS AND ACRONYMS 

ACC American Chemistry Council 

ADI Acceptable Daily Intake 

AEGL acute exposure guideline level 

ARAR applicable or relevant and appropriate requirement 

ARE acute reference exposure 

AQI Air Quality Index 

ATFERM Agency Task Force on Environmental Regulatory Modeling 

ATSDR Agency for Toxic Substances and Disease Registry 

AUC area under the curve 

BAF bioaccumulation factor 

BAT best available technology 

BMC benchmark concentration 

BMCL benchmark concentration lower confidence limit 

BMD benchmark dose 

BMDL benchmark dose lower confidence level 

BMR benchmark response 

CAA Clean Air Act 

CAAA Clean Air Act Amendments 

CBD chronic beryllium disease 

CDC Centers for Disease Control and Prevention 

CEPPO Chemical Emergency Preparedness and Prevention Office 

CERCLA Comprehensive Environmental Response, Compensation, and Liability 

Act 

CERCLIS Comprehensive Environmental Response, Compensation, and Liability 

Information System 

CFSAN Center for Food Safety and Nutrition 

CNS central nervous system 

CPSC Consumer Product Safety Commission 

CRARM Congressional/Presidential Commission on Risk Assessment and Risk 

Management 

CREM Committee on Regulatory Environmental Modeling 

CSAF chemical-specific adjustment factor 

CSF cancer slope factor 

CSFII Continuing Survey of Food Intakes by Individuals 

CTE central tendency exposure 

CWA Clean Water Act 

DAF dosimetric adjustment factor 

DBP disinfection byproduct 

DENR Department of Environment and Natural Resources 

DHHS Department of Health and Human Services 
Examination of EPA Risk Assessment Principles and Practices Page 149 

DHS Department of Homeland Security 

DNA deoxyribonucleic acid 

DNT developmental neurotoxicity 

DOD Department of Defense 

DOE Department of Energy 

DOJ Department of Justice 

DOT Department of Transportation 

ECC Environmental Clearance Committee 

Eco-SSLs Ecological Soil Screening Levels 

ED effective dose 

EDC endocrine disrupting chemical 

EMAP Environmental Monitoring and Assessment Program 

EPA Environmental Protection Agency 

EPCRA Emergency Planning and Community Right-To-Know Act 

ETV Environmental Technology Verification 

FAV Final Acute Value 

FDA Food and Drug Administration 

FDAMA Food and Drug Administration Modernization Act 

FFDCA Federal Food, Drug, and Cosmetic Act 

FIELDS Field Environmental Decision Support 

FIFRA Federal Insecticide, Fungicide, and Rodenticide Act 

FQPA Food Quality Protection Act 

FR Federal Register 

FS feasibility study 

GAO General Accounting Office 

GHS globally harmonized system 

GIS geographic information system 

GLI Great Lakes Initiative 

GLNPO Great Lakes National Program Office 

GPRA Government Performance and Results Act 

GPS Global Positioning System 

GSA General Services Administration 

HA Health Advisory 

HAP hazardous air pollutant 

HAZMAT hazardous materials 

HEC human equivalent concentration 

HED human equivalent dose 

HEPA high-efficiency particulate air 

HI hazard index 

HMIS Hazardous Materials Information System 

HQ hazard quotient 

HVAC heating, ventilation, and air conditioning 
Page 150 Examination of EPA Risk Assessment Principles and Practices 

IBI Index of Biological Integrity 

IC index compound 

ICRP International Commission on Radiological Protection 

IEUBK Integrated Exposure Uptake Biokinetic Model for Lead in Children 

IME individuals most exposed 

IPCS International Programme on Chemical Safety 

IRIS Integrated Risk Information System 

ISO International Organization for Standardization 

LED lowest effective dose 

LOAEL lowest-observed-adverse-effect level 

LOEL lowest-observed-effect level 

MACT maximum achievable control technology 

MCL maximum contaminant level 

MCLG maximum contaminant level goal 

MEI maximally exposed individual 

MF modifying factor 

MIMS Multimedia Integrated Modeling System 

MLE maximum likelihood estimate 

MOA mode of action 

MOE margin of exposure 

MOU Memorandum of Understanding 

MRL minimal risk level 

MTD maximum tolerated dose 

NAAQS National Ambient Air Quality Standards 

NAS National Academy of Sciences 

NATO/CCMS North Atlantic Treaty Organization Committee on the Challenges of 

Modern Society 

NCEA National Center for Environmental Assessment 

NCER National Center for Environmental Research 

NCP National Oil and Hazardous Substances Pollution Contingency Plan 

NCTR National Center for Toxicological Research 

NEIC National Enforcement Investigation Center 

NERL National Environmental Research Laboratory 

NESHAP National Emissions Standard for Hazardous Air Pollutants 

NHEERL National Health and Environmental Effects Research Laboratory 

NHSRC National Homeland Security Research Center 

NIEHS National Institute for Environmental Health Sciences 

NIOSH National Institute for Occupational Safety and Health 

NNI National Nanotechnology Initiative 

NOAA National Oceanic and Atmospheric Administration 

NOAEL no-observed-adverse-effect level 

NOEL no-observed-effect level 
Examination of EPA Risk Assessment Principles and Practices Page 151 

NPDES National Pollutant Discharge Elimination System 

NRC National Research Council 

NRMRL National Risk Management Research Laboratory 

NSF National Science Foundation 

NTP National Toxicology Program 

OAQPS Office of Air Quality, Planning, and Standards 

OAR Office of Air and Radiation 

OECD Organisation for Economic Cooperation and Development 

OERR Office of Emergency and Remedial Response 

OMB Office of Management and Budget 

OPEI Office of Policy, Economics, and Innovation 

OPP Office of Pesticide Programs 

OPPT Office of Pollution Prevention and Toxics 

OPPTS Office of Prevention, Pesticides and Toxic Substances 

ORD Office of Research and Development 

OSHA Occupational Safety and Health Administration 

OSTP Office of Science and Technology Policy 

OSW Office of Solid Waste 

OSWER Office of Solid Waste and Emergency Response 

OU operable unit 

OW Office of Water 

OWOW Office of Wetlands, Oceans, and Watersheds 

PAD population-adjusted dose 

PAH polycyclic aromatic hydrocarbon 

PBPK physiologically based pharmacokinetic 

PCB polychlorinated biphenyl 

PCDD polychlorinated dibenzo-p-dioxin 

PCDF polychlorinated dibenzofurans 

PEL permissible exposure limit 

PM particulate matter 

PMN premanufacture notification 

POD point of departure 

PPE personal protective equipment 

ppm parts per million 

PRG Preliminary Remediation Goal 

QA quality assurance 

QAPP Quality Assurance Project Plan 

QA/QC quality assurance/quality control 

QSAR quantitative structure-activity relationship 

RAF Risk Assessment Forum 

RAGS Risk Assessment Guidance for Superfund 

RCRA Resource Conservation and Recovery Act 
Page 152 Examination of EPA Risk Assessment Principles and Practices 

RDDR regional deposited dose ratio 

ReVA Regional Vulnerability Assessment 

RGDR regional gas dose ratio 

RfC reference concentration 

RfD reference dose 

RI remedial investigation 

RME reasonable maximum exposure 

RNA ribonucleic acid 

ROD Record of Decision 

RPF relative potency factor 

RQ risk quotient 

RSC relative source contribution 

SAB Science Advisory Board 

SAP Scientific Advisory Panel 

SAR structure-activity relationship 

SAV Secondary Acute Value 

SARA Superfund Amendments and Reauthorization Act 

SCV Secondary Chronic Value 

SDWA Safe Drinking Water Act 

SDWAA Safe Drinking Water Act Amendments 

SEER Surveillance, Epidemiology, and End Results 

SEQL Sustainable Environment for Quality of Life 

SIP State Implementation Plan 

SMART simple multi-attribute rating technique 

SMCL secondary maximum contaminant level 

SOPs standard operating procedures 

STEL short-term exposure limit 

STAR Science To Achieve Results 

TCCR transparency, clarity, consistency, reasonableness 

TCDD tetrachlorodibenzo-p-dioxin 

TEF Toxicity Equivalency Factor 

TEQ Toxic Equivalency Quotient 

TI tolerable intake 

TIO Technology Innovation Office 

TRC Toxicogenomics Research Consortium 

TRI Toxics Release Inventory 

TRV toxicity reference value 

TSCA Toxic Substances Control Act 

TWA time-weighted average 

UCL upper confidence limit 

UCL95 95% upper confidence limit of the arithmetic mean 
Examination of EPA Risk Assessment Principles and Practices Page 153 

UF uncertainty factor 

USDA United States Department of Agriculture 

WHO-ECEH World Health Organization European Centre for Environmental Health 

WQC water quality criteria 
Page 154 Examination of EPA Risk Assessment Principles and Practices 

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U.S. Environmental Protection Agency (USEPA). (1993d). Water quality guidance for the great 

lakes system, Proposed rule. Fed Reg 58:20802.ff April 16. 

U.S. Environmental Protection Agency (USEPA). (1994a). Methods for Derivation of Inhalation 

reference concentrations and application of inhalation dosimetry. EPA/600/8-90/066F. 

Office of Research and Development, Washington, DC. 

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environmental regulatory modeling — Guidance, support needs, draft criteria and charter. 

EPA 500-R-94-001. Washington, DC. 

U.S. Environmental Protection Agency (USEPA). (1994c). Model validation for predictive 

exposure assessments. Risk Assessment Forum, Washington, DC. 

(http://cfpub.epa.gov/crem/cremlib.cfm) 
Page 168 Examination of EPA Risk Assessment Principles and Practices 

U.S. Environmental Protection Agency (USEPA). (1994d). Peer review program. Memorandum 

from Administrator Carol Browner to Assistant Administrators, General Counsel, 

Inspector General, Associate Administrators, Regional Administrators, and Staff Office 

Directors. Science Policy Council, Washington, DC. June 7. 

(http://www.epa.gov/osp/spc/memo0607.htm) 

U.S. Environmental Protection Agency (USEPA). (1994e). Catalog of standard toxicity tests for 

ecological risk assessment. EPA 540-F-94-013. Office of Emergency and Remedial 

Response, Washington, DC. 

U.S. Environmental Protection Agency (USEPA). (1994f). Field studies for ecological risk 

assessment. EPA 540-F-94-014. Office of Emergency and Remedial Response, 

Washington, DC. 

U.S. Environmental Protection Agency (USEPA). (1994g). A review of ecological assessment 

case studies from a risk assessment perspective, Vol. II. EPA 630-R-94-003. Washington, 

D.C. 

U.S. Environmental Protection Agency (USEPA). (1994h). Ecological risk assessment issue 

papers. EPA/630/R-94/009. Risk Assessment Forum, Washington, DC. 

U.S. Environmental Protection Agency (USEPA). (1994i). Managing ecological risks at EPA: 

Issues and recommendations for progress. EPA/600/R-94/183. Washington, DC. 

U.S. Environmental Protection Agency (USEPA). (1995a). Policy for risk characterization. 

Science Policy Council, Washington, DC. (http://www.epa.gov/osp/spc/rcpolicy.htm) 

U.S. Environmental Protection Agency (USEPA). (1995b). Policy on evaluating health risks to 

children. Science Policy Council, Washington, DC. 

(http://www.epa.gov/osp/spc/memohlth.htm) 

U.S. Environmental Protection Agency (USEPA). (1995c). Water quality guidance for the Great 

Lakes system, Final Rule. Fed Reg 60(56): 15366-15425. 

U.S. Environmental Protection Agency (USEPA). (1995d). Water quality guidance for the Great 

Lakes system: Supplementary information document (SID). EPA-820-B-95-001. Office 

of Water, Washington, DC. 

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guide. 2nd ed. OSWER Pub. 9355.4-23. Office of Emergency and Remedial Response, 

Washington, DC. 
Examination of EPA Risk Assessment Principles and Practices Page 169 

U.S. Environmental Protection Agency (USEPA). (1996b). Recommendations of the technical 

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exposures to lead in soil. EPA-540-R-03-001. Washington, DC. 

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(http://www.epa.gov/superfund/programs/reforms/reforms/3-5a.htm#des) 

U.S. Environmental Protection Agency (USEPA). (1996d). PCBs: Cancer dose-response 

assessment and application to environmental mixtures. EPA/600/P-96/001F. National 

Center for Environmental Assessment, Washington, DC. 

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assessment. Part 1: Planning and scoping. Science Policy Council, Washington, DC. 

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U.S. Environmental Protection Agency (USEPA). (1997b). Guiding principles for Monte Carlo 

analysis. EPA/630/R-97/001. Risk Assessment Forum, Office of Research and 

Development, Washington, DC. 

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analysis in risk assessment at the U.S. Environmental Protection Agency. Fred Hansen, 

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9285.7-25, EPA 540-R-97-006, PB97-963211, Washington, DC. 

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assessment. EPA/630/R-95/002F. Risk Assessment Forum, Washington, DC. 171 pp. 

(http://cfpub.epa.gov/ncea/cfm/recordisplay.cfm?deid=12460) 
Page 170 Examination of EPA Risk Assessment Principles and Practices 

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(http://www.epa.gov/oscpmont/sap/1998/index.htm) 

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risks from pesticides and their uses: Background, methodology & case study. Office of 

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(http://www.epa.gov/scipoly/sap/1998/index.htm#december8) 

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proposed plans, records of decision, and other remedy selection decision documents. EPA 

540/R-98/031. Office of Solid Waste and Emergency Response, Washington, DC. 

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assessment, Revised draft. NCEA-F-0644. Office of Research and Development, 

Washington, DC. 

(http://cfpub.epa.gov/ncea/raf/cancer_gls.pdf) 

U.S. Environmental Protection Agency (USEPA). (1999c). White paper on the nature and scope 

of issues on adoption of model use acceptability guidance. Science Policy Council., 

Washington, DC. 

(http://cfpub.epa.gov/crem/cremlib.cfm) 

U.S. Environmental Protection Agency (USEPA). (1999d). Proposed methods for determining 

watershed-derived percent crop areas and considerations for applying crop area 

adjustments to surface water screening models. FIFRA Scientific Advisory Panel, 

Washington, DC. May 25-27. 

(http://www.epa.gov/oscpmont/sap/1999/index.htm) 
Examination of EPA Risk Assessment Principles and Practices Page 171 

U.S. Environmental Protection Agency (USEPA). (1999e). Two documents: 1) Downwind 

deposition tolerance bounds for ground hydraulic boom sprayers; 2) Downwind 

deposition tolerance bounds for orchards. FIFRA Scientific Advisory Panel, Washington, 

DC. July 20-23. 

(http://www.epa.gov/scipoly/sap/1999/index.htm) 

U.S. Environmental Protection Agency (USEPA). (1999f). Risk assessment guidance for 

Superfund, Volume 1, Human health evaluation manual (Supplement to Part A): 

Community involvement in Superfund risk assessments. EPA 540-R-98-042. Office of 

Solid Waste and Emergency Response, Washington, DC. March. 

U.S. Environmental Protection Agency (USEPA). (1999g). Short Sheet: IEUBK Model 

Bioavailability Variable. EPA 540-F-00-006. Office of Solid Waste and Emergency 

Response, Washington, DC. October. 

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you can help (videotape). EPA-540-V-99-002. Office of Solid Waste and Emergency 

Response, Washington, DC. September. 

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in streams and wadeable rivers: Periphyton, benthic macroinvertebrates and fish, 2nd 

edition. EPA 841-B-99-002. Office of Water, Washington, DC. 

U.S. Environmental Protection Agency (USEPA). (1999j). Issuance of final guidance: Ecological 

risk assessment and risk management principles for Superfund Sites. OSWER Directive 

9285.7-28 P. Office of Solid Waste and Emergency Response, Washington, DC. October 

7. 

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Risk Characterization Handbook. EPA 100-B00-002. Science Policy Council, 

Washington, DC. December. (http://www.epa.gov/osp/spc/rchandbk.pdf) 

U.S. Environmental Protection Agency (USEPA). (2000b). Science Policy Council Handbook: 

Peer Review Handbook, 2nd edition. EPA 100-B00-001. Science Policy Council, 

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U.S. Environmental Protection Agency (USEPA). (2000c). Policy and program requirements for 

the mandatory agency-wide quality system. EPA Order 5360.1 A2. Office of 

Environmental Information, Washington, DC. 

(http://www.epa.gov/quality/qs-docs/5360-1.pdf) 
Page 172 Examination of EPA Risk Assessment Principles and Practices 

U.S. Environmental Protection Agency (USEPA). (2000d). EPA quality manual for 

environmental programs. EPA Order 5360 A1. Office of Environmental Information, 

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U.S. Environmental Protection Agency (USEPA). (2000e). An SAB advisory on the USEPA’s 

draft case study analysis of the residual risk of secondary lead smelters, EPA Science 

Advisory Board, May 2000, pp. 20-21. 

(http://www.epa.gov/sab/pdf/ecadv05.pdf) 

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consumption in the United States. EPA-821-R-00-025. 

U.S. Environmental Protection Agency (USEPA). (2000g). Benchmark dose technical guidance 

document. EPA/630/R-00/001. Risk Assessment Forum, Washington, DC. 

(http://cfpub1.epa.gov/ncea/cfm/recordisplay.cfm?deid=22506) 

U.S. Environmental Protection Agency (USEPA). (2000h). Supplementary guidance for 

conducting health risk assessment of chemical mixtures. EPA/630/R-00/002. Risk 

Assessment Forum, Washington DC. 

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reassessment of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) and related compounds 

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Assessment, Office of Research and Development, Washington, DC. 

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estimating pesticide concentrations in drinking water and assessing water treatment 

effects on pesticide removal and transformation: A consultation. FIFRA Scientific 

Advisory Panel, Washington, DC. September, 26-29. 

(http://www.epa.gov/oscpmont/sap/2000/index.htm) 

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water quality criteria for the protection of human health. EPA-822-B-00-004. Office of 

Water, Office of Science and Technology, Washington, DC. 

U.S. Environmental Protection Agency (USEPA). (2000l). Data quality objectives process for 

hazardous waste site investigations — Data quality objectives guidance. EPA/600/R00/007. Office of Environmental Information, Washington, DC. January. 
Examination of EPA Risk Assessment Principles and Practices Page 173 

U.S. Environmental Protection Agency (USEPA). (2000m). Phase 2 report: Further site 

characterization and analysis, Volume 2F. Revised human health risk assessment Hudson 

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U.S. Environmental Protection Agency (USEPA). (2000n). Superfund community involvement 

guidance. EPA 540-K-01-003. Office of Solid Waste and Emergency Response, 

Washington, DC. April. 

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EPA 540-R-02-002. Office of Emergency and Remedial Response, Washington, DC. 

December. 

(http://www.epa.gov/superfund/programs/risk/rags3a/index.htm) 

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Superfund, Volume I, Human health evaluation manual (Part D, Standardized planning, 

reporting and review of Superfund risk assessments) Final. Office of Solid Waste and 

Emergency Response, Washington, DC. December. 

(http://www.epa.gov/superfund/programs/risk/ragsd) 

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Information System, Office of Research and Development, Washington, DC. 

(http://www.epa.gov/iris/subst/0025.htm) 

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model for lead in children (IEUBKwin v1.0). Washington, DC. 

(http://www.epa.gov/superfund/programs/lead/ieubk.htm) 

U.S. Environmental Protection Agency (USEPA). (2001e). Workshop report on the application 

of 2,3,7,8-TCDD toxicity equivalence factors to fish and wildlife. EPA/630/R-01/002. 

Risk Assessment Forum, Washington, DC. 

(http://cfpub.epa.gov/ncea/raf/recordisplay.cfm?deid=23763) 

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assurance project plans, EPA QA/R-5. EPA/240/B-01/003. Office of Environmental 

Information, Washington, DC. 

(http://www.epa.gov/QUALITY/qs-docs/r5-final.pdf) 
Page 174 Examination of EPA Risk Assessment Principles and Practices 

U.S. Environmental Protection Agency (USEPA). (2001g). Risk assessment guidance for 

Superfund, Volume I, Human health evaluation manual (Part E): Supplemental guidance 

for dermal risk assessment, Interim. PB99-963312. Office of Solid Waste and Emergency 

Response, Washington, DC. May. 

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Comfort)/Lavaca Bay Site Point Comfort, TX. CERCLIS #TXD008123168. Region VI, 

Dallas, TX. December. 

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guidance. EPA 40-R-01-007. Office of Emergency and Remedial Response, Washington, 

DC. June. 

U.S. Environmental Protection Agency (USEPA). (2001j). National-scale air toxics assessment 

for 1996. Science Advisory Board Preliminary Draft. EPA-453/R-01-003, Office of Air 

Quality Standards and Planning, Research Triangle Park, NC. 

(http://www.epa.gov/ttn/atw/nata/natsaov.html) 

U.S. Environmental Protection Agency (USEPA). (2002a). Guidelines for ensuring and 

maximizing the quality, objectivity, utility, and integrity, of information disseminated by 

the Environmental Protection Agency. EPA/260R-02-008. Office of Environmental 

Information, Washington, DC. October. 

(http://www.epa.gov/oei/qualityguidelines) 

U.S. Environmental Protection Agency (USEPA). (2002b). Lessons learned on planning and 

scoping for environmental risk assessments. Science Policy Council. January. 

(http://www.epa.gov/osp/spc/2cumrisk.htm) 

U. S. Environmental Protection Agency (USEPA). (2002c). A review of the reference dose and 

reference concentration processes. EPA/630/P-02/002F. Risk Assessment Forum, 

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U.S. Environmental Protection Agency (USEPA). (2002d). Child-specific exposure factors 

handbook (Interim report). EPA-600-P-00-002B. Office of Research and Development, 

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Examination of EPA Risk Assessment Principles and Practices Page 175 

U.S. Environmental Protection Agency (USEPA). (2002f). Guidance on cumulative risk 

assessment of pesticide chemicals that have a common mechanism of toxicity. Office of 

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TRIM.FaTE technical support document, Volume I: Description of module. EPA-453/R02-011a. Office of Air Quality Planning and Standards, Research Triangle Park, NC. 

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TRIM.FaTE technical support document, Volume II: Description of chemical transport 

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Region III, Philadelphia, PA. Approach and initial findings. EPA-453/R-02-012. Office 

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developing soil screening levels for Superfund sites. OSWER 9355.4-24. Office of Solid 

Waste and Emergency Response, Washington, DC. 

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for exposure point concentrations at hazardous waste sites. OSWER 9285.6-10. Office of 

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Pile, Palmerton, PA. EPD ID: PAD002395887.OU 03. EPA/ROD/R03-02/007. 

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approaches to polycyclic aromatic hydrocarbon (PAH) health assessment. EPA/635/R02/005. Office of Research and Development, Washington, DC. 

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Page 176 Examination of EPA Risk Assessment Principles and Practices 

U.S. Environmental Protection Agency (USEPA). (2003b). Framework for cumulative risk 

assessment. EPA/630/P-02/001F. Office of Research and Development, Washington, DC. 

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risk assessment. EPA/630/P-03/001A. NCEA-F-0644A. Risk Assessment Forum, 

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U.S. Environmental Protection Agency (USEPA). (2003d). Supplemental guidance for assessing 

cancer susceptibility from early-life exposure to carcinogens (external review draft). 

EPA/630/R-03/003. Risk Assessment Forum, Washington, DC. 

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for pesticide mixtures: Biostatistical analyses of joint dose-response. EPA/600/R-03/052. 

Office of Research and Development, National Center for Environmental Assessment, 

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(http://www.epa.gov/iris) 

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hazardous air pollutants for stationary combustion turbines; Proposed rule. 40 CFR Part 

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U.S. Environmental Protection Agency (USEPA). (2003j). Generic endpoints for ecological risk 

assessment. EPA/630/P-02/004A. Risk Assessment Forum, Washington, DC. 

U.S. Environmental Protection Agency (USEPA). (2003k). Guidance for developing ecological 

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Examination of EPA Risk Assessment Principles and Practices Page 177 

U.S. Environmental Protection Agency (USEPA). (2003l). Potential risks of nine rodenticides to 

birds and nontarget mammals: A comparative approach. January. 

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Page 178 Examination of EPA Risk Assessment Principles and Practices 

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Page 180 Examination of EPA Risk Assessment Principles and Practices 

REFERENCES OF EPA RISK ASSESSMENT GUIDELINES 

Guidelines for Carcinogen Risk Assessment. Federal Register 51: 33992-34003, 24 September 

1986; also EPA Publication No. EPA/600/8-87/045, August 1987. 

Proposed Guidelines for Carcinogen Risk Assessment; Notice. Federal Register 61: 

17960-18011, 23 April 1996. 

Draft Revised Guidelines for Carcinogen Risk Assessment. EPA Publication No. NCEA-F-0644, 

July 1999; Notice in Federal Register 66: 59593-59594, 29 November 2001. 

Draft Final Guidelines for Carcinogen Risk Assessment (External Review Draft, February 2003). 

EPA Publication No. NCEA-F-0644A, March 2003; Notice in Federal Register 68: 

10012-10015, 3 March 2003. 

Supplemental Guidance for Assessing Cancer Susceptibility from Early-life Exposure to 

Carcinogens (External Review Draft). EPA Publication No. EPA/630/R-03/003, February 

2003; Notice in Federal Register 68: 10012-10015, 3 March 2003. 

Guidelines for Mutagenicity Risk Assessment. Federal Register 51: 34006-34012, 24 September 

1986; also EPA Publication No. EPA/600/8-87/045, August 1987. 

Guidelines for the Health Risk Assessment of Chemical Mixtures. Federal Register 51: 

34014-34025, 24 September 1986; also EPA Publication No. EPA/600/8-87/045, August 

1987. 

Supplementary Guidance for Conducting Health Risk Assessment of Chemical Mixtures. EPA 

Publication No. EPA/630/R-00/002, August 2000. 

Guidelines for Developmental Toxicity Risk Assessment. Federal Register 56: 63798-63826, 5 

December 1991. 

Guidelines for Exposure Assessment. Federal Register 57: 22888-22938, 29 May 1992; also EPA 

Publication No. EPA/600/Z-92/001. 

Guidelines for Reproductive Toxicity Risk Assessment; Notice. Federal Register 61: 

56274-56322, 31 October 1996. 

Assessment of Thyroid Follicular Cell Tumors. EPA Publication No. EPA/630/R-97/002, March 

1998. 
Examination of EPA Risk Assessment Principles and Practices Page 181 

Guidelines for Ecological Risk Assessment. Federal Register 63: 26846-26924, 14 May 1998; 

also EPA Publication No. EPA/630/R-95/002F, April 1998. 

Guidelines for Neurotoxicity Risk Assessment; Notice. Federal Register 60: 26926-26954, 14 

May 1998. 

Guiding Principles for Monte Carlo Analysis (contains: Policy for Use of Probabilistic Analysis 

in Risk Assessment at the U.S. Environmental Protection Agency). EPA Publication No. 

EPA/630/R-97/001, March 1997. 
Page 182 Examination of EPA Risk Assessment Principles and Practices 

ADDITIONAL USEFUL WEB SITES 

EPA Risk Assessment Guidelines Web site: 

http://cfpub.epa.gov/ncea/raf/recordisplay.cfm?deid=55907 

EPA Quality System Web site: http://www.epa.gov/quality 

EPA Science Policy Council Web site: http://www.epa.gov/osp/spc 

EPA Information Quality Guidelines Web site: http://www.epa.gov/oei/qualityguidelines