EPA 120/R-07/001 | March 2007 

www.epa.gov/osa 

Framework for 

Metals Risk Assessment 

Office of the Science Advisor 

Risk Assessment Forum 
 EPA 120/R-07/001 

March, 2007 

Framework for 

Metals Risk Assessment 

Office of the Science Advisor 

Risk Assessment Forum 

U.S. Environmental Protection Agency 

Washington, DC 20460
DISCLAIMER 

This document has been reviewed in accordance with U.S. Environmental Protection 

Agency policy. Mention of trade names or commercial products does not constitute endorsement 

or recommendation for use. 

ii 
CONTENTS 

LIST OF TABLES....................................................................................................................... vii 

LIST OF FIGURES ..................................................................................................................... vii 

LIST OF ABBREVIATIONS AND ACRONYMS ...................................................................viii 

PREFACE...................................................................................................................................xiii 

AUTHORS.................................................................................................................................. xvi 

EXECUTIVE SUMMARY ....................................................................................................... xvii 

1. INTRODUCTION .................................................................................................................1-1 

1.1. PURPOSE AND AUDIENCE ................................................................................1-1 

1.2. METALS FRAMEWORK SCOPE.........................................................................1-3 

1.3. METALS ASSESSMENT CONTEXT...................................................................1-3 

1.3.1. National Ranking and Categorization.........................................................1-4 

1.3.2. National-Level Assessments.......................................................................1-6 

1.3.3. Site-Specific Assessments ..........................................................................1-8 

1.4. KEY PRINCIPLES TO CONSIDER IN METALS RISK ASSESSMENT...........1-9 

1.4.1. Metals are Naturally Occurring Constituents in the Environment and 

Vary in Concentrations Across Geographic Regions................................1-9 

1.4.2. All Environmental Media have Naturally Occurring Mixtures of 

Metals and Metals are Often Introduced into the Environment as 

Mixtures...................................................................................................1-10 

1.4.3. Some Metals are Essential for Maintaining Proper Health of Humans, 

Animals, Plants, and Microorganisms.....................................................1-10 

1.4.4. The Environmental Chemistry of Metals Strongly Influences Their 

Fate and Effects on Human and Ecological Receptors............................1-10 

1.4.5. The Toxicokinetics and Toxicodynamics of Metals Depend on the 

Metal, the Form of the Metal or Metal Compound, and the 

Organism’s Ability to Regulate and/or Store the Metal ..........................1-11

2. FRAMEWORK FOR METALS RISK ASSESSMENT.......................................................2-1 

2.1. HUMAN HEALTH AND ECOLOGICAL RISK ASSESSMENT: 

PLANNING AND PROBLEM FORMULATION ................................................2-1 

2.2. METALS CONCEPTUAL MODEL ......................................................................2-4 

2.3. ASSESSMENT PHASE........................................................................................2-10 

2.3.1. Bioavailability...........................................................................................2-10 

2.3.2. Exposure Characterization........................................................................2-11 

2.3.3. Characterization of Effects/Hazard Analysis............................................2-14 

2.4. RISK CHARACTERIZATION ............................................................................2-16

3. ENVIRONMENTAL CHEMISTRY, TRANSPORT AND FATE.......................................3-1 

3.1. INTRODUCTION AND TERMINOLOGY...........................................................3-1 

3.1.1. Hard and Soft Acids and Bases: The Stability of Complexes ...................3-2 

3.1.2. Transformations..........................................................................................3-3 

3.1.2.1. Biotic Methylation Transformations............................................3-5 

3.1.2.2. Abiotic Transformations ..............................................................3-6 

3.1.3. Aquatic Chemistry ......................................................................................3-7 

3.1.3.1. Processes .....................................................................................3-7 

3.1.3.2. Speciation and Complexes...........................................................3-7 

iii 
3.1.3.3. Importance of pH and Redox Conditions ....................................3-9 

3.1.4. Ground Water and Metals Mobility............................................................3-9 

3.1.4.1. Application of Partition Coefficients to Metal Mobility in 

Ground Water............................................................................3-11 

3.1.5. Sediment Chemistry..................................................................................3-13 

3.1.5.1. Metal Chemistry in Sediments ...................................................3-14 

3.1.5.2. Estimating Metal Distribution in Sediments..............................3-16 

3.1.6. Soil Chemistry ..........................................................................................3-17 

3.1.6.1. Key Parameters Affecting Metal Bioavailability in Soils ..........3-18 

3.1.6.2. Adsorption Behavior of the Metals of Concern .........................3-19 

3.1.6.3. Aging of Metals in Soil ..............................................................3-19 

3.1.6.4. Dissolution and Transformation of Metals................................3-20 

3.1.6.5. Soil Metal Transfer to Plants.....................................................3-20 

3.1.7. Atmospheric Behavior/Chemistry ............................................................3-21 

3.2. METAL FATE AND TRANSPORT ....................................................................3-23 

3.2.1. Aquatic and Terrestrial Transport Pathways for Metals...........................3-24 

3.2.1.1. Aquatic Transport Models .........................................................3-26 

3.2.2. Terrestrial Fate and Transport...................................................................3-29 

3.3. ATMOSPHERIC FATE AND TRANSPORT......................................................3-29

4. HUMAN HEALTH RISK ASSESSMENT FOR METALS.................................................4-1 

4.1. METALS PRINCIPLES..........................................................................................4-2 

4.2. HUMAN EXPOSURE ASSESSMENT..................................................................4-3 

4.2.1. Background Levels.....................................................................................4-3 

4.2.2. Bioavailability.............................................................................................4-4 

4.2.3. Susceptible Populations..............................................................................4-5 

4.2.3.1. Life Stage .....................................................................................4-6 

4.2.3.2. Demographics..............................................................................4-6 

4.2.3.3. Pregnancy and Lactation.............................................................4-6 

4.2.3.4. Concurrent Damage or Disease ..................................................4-6 

4.2.3.5. Nutritional State...........................................................................4-7 

4.2.3.6. Genetic Polymorphisms and Variability......................................4-7 

4.2.4. Environmental Release, Transport, and Fate ..............................................4-8 

4.2.5. Route-Specific Differences in Effects ........................................................4-8 

4.2.5.1. Inhalation Exposure.....................................................................4-8 

4.2.5.2. Dietary Exposure .......................................................................4-10 

4.2.5.3. Incidental Soil Exposure............................................................4-11 

4.2.5.4. Drinking Water Exposure ..........................................................4-11 

4.2.5.5. Dermal Exposure .......................................................................4-12 

4.2.6. Integrated Exposure ..................................................................................4-12 

4.2.6.1. Modeling ....................................................................................4-12 

4.2.7. Biomarkers................................................................................................4-13 

4.3. HAZARD CHARACTERIZATION.....................................................................4-14 

4.3.1. Mixtures and Interactions .........................................................................4-14 

4.3.2. Essentiality................................................................................................4-16 

4.3.3. Forms of Metals........................................................................................4-18 

4.3.4. Toxicokinetics/Toxicodynamics...............................................................4-18 

4.3.4.1. Absorption..................................................................................4-20 

4.3.4.2. Distribution................................................................................4-20 

iv 
4.3.4.3. Metabolism.................................................................................4-21 

4.3.4.4. Excretion....................................................................................4-21 

4.3.4.5. Kinetic Modeling........................................................................4-22 

4.3.5. Metal Toxicity...........................................................................................4-23 

4.3.5.1. Noncancer Effects of Metals......................................................4-23 

4.3.5.2. Carcinogenic Effects of Metals..................................................4-24 

4.3.5.3. Issues Related to Evaluation of Toxicity Tests for Metals.........4-25 

4.3.6. Dose-Response Assessment......................................................................4-26 

4.4. RISK CHARACTERIZATION ............................................................................4-27

5. AQUATIC ECOLOGICAL RISK ASSESSMENT FOR METALS.....................................5-1 

5.1. METALS PRINCIPLES..........................................................................................5-1 

5.2. CHARACTERIZATION OF EXPOSURE.............................................................5-2 

5.2.1. Background Levels.....................................................................................5-2 

5.2.2. Forms of Metals..........................................................................................5-3 

5.2.3. Exposure Pathway Analysis........................................................................5-4 

5.2.4. Fate and Transport of Metals......................................................................5-5 

5.2.5. Toxicokinetics and Toxicodynamics (Bioavailability and 

Bioaccumulation Issues)............................................................................5-6 

5.2.5.1. Aqueous Phase.............................................................................5-6 

5.2.5.2. Sediment Phase............................................................................5-8 

5.2.5.3. Dietary Phase ..............................................................................5-9 

5.2.5.4. Bioaccumulation and Trophic Transfer.....................................5-11 

5.3. CHARACTERIZATION OF EFFECTS...............................................................5-14 

5.3.1. Essentiality................................................................................................5-14 

5.3.2. Toxicokinetics/Toxicodynamics (Toxicity Issues)...................................5-14 

5.3.3. Metal Mixtures..........................................................................................5-15 

5.3.4. Critical Body Residues .............................................................................5-17 

5.3. RISK CHARACTERIZATION ............................................................................5-19

6. TERRESTRIAL ECOLOGICAL RISK ASSESSMENT FOR METALS............................6-1 

6.1. METALS PRINCIPLES..........................................................................................6-1 

6.2. CHARACTERIZATION OF EXPOSURE.............................................................6-1 

6.2.1. Natural Occurrence of Metals.....................................................................6-1 

6.2.2. Forms of Metals..........................................................................................6-3 

6.2.3. Exposure Routes.........................................................................................6-3 

6.2.4. Soil Transport and Fate Models..................................................................6-3 

6.2.5. Toxicokinetics/Toxicodynamics.................................................................6-4 

6.2.5.1. Bioavailability..............................................................................6-4 

6.2.5.2. Bioaccumulation ..........................................................................6-5 

6.2.6. Soil Invertebrate Exposure..........................................................................6-6 

6.2.7. Plant Exposure ............................................................................................6-7 

6.2.8. Wildlife Exposure .......................................................................................6-8 

6.3. CHARACTERIZATION OF EFFECTS...............................................................6-11 

6.3.1. Essentiality................................................................................................6-11 

6.3.2. Toxicity Tests ...........................................................................................6-13 

6.3.3. Metal Mixtures..........................................................................................6-14 

6.3.4. Critical Body Residues .............................................................................6-15 

6.3.5. Plant and Soil Invertebrate Toxicity .........................................................6-16 

v 
6.3.6. Wildlife Toxicity.......................................................................................6-17 

6.4. RISK CHARACTERIZATION ............................................................................6-18

REFERENCES .......................................................................................................................... R-1

APPENDIX A: FRAMEWORK FOR METALS RISK ASSESSMENT

CONTRIBUTORS AND REVIEWERS........................................................................ A-1 

vi 
LIST OF TABLES 

Table 3-1. Hard and soft acids (metal cations) and bases (ligands)............................................ 3-3 

Table 3-2. Examples of organometallic compounds................................................................... 3-4 

Table 3-3. Metals/metalloids involved in methylation processes............................................... 3-5 

Table 3-4. General trends of environmental factors affecting rates of

methylation/demethylation ........................................................................................ 3-6 

Table 4-1. Metal essentiality for humans.................................................................................. 4-17

Table 4-2. Summary of major differences in kinetic behavior of organic compounds 

compared to metals and inorganic metal compounds in humans ............................ 4-19 

Table 6-1. Qualitative bioavailability of metal cations in natural soils to plants and soil 

invertebrates............................................................................................................... 6-8

Table 6-2. Qualitative bioavailability of metal anions in natural soils to plants and soil 

invertebrates............................................................................................................... 6-8

Table 6-3. Metals classified by their known essentiality.......................................................... 6-13 

LIST OF FIGURES 

Figure 1-1. Categories of metals assessments under EPA statutory framework ........................ 1-5 

Figure 2-1. Risk assessment/risk management process for metals............................................. 2-2 

Figure 2-2. Generic conceptual model for metals risk assessment............................................. 2-6 

Figure 2-3. Conceptual diagram for evaluating bioavailability processes and 

bioaccessibility for metals in soil, sediment, or aquatic systems ........................... 2-12 

Figure 3-1. Approximate positions of some natural environments in terms of Eh and pH ...... 3-10 

Figure 3-2. A generalized model framework for chemical fate and transport in an 

 aquatic system......................................................................................................... 3-26 

Figure 6-1. Generalized representation of percent contribution of incidental soil 

ingestion to oral dose for wildlife at different soil ingestion rates and 

bioaccumulation factors and a bioavailability of 100 percent ................................ 6-10 

vii 
LIST OF ABBREVIATIONS AND ACRONYMS 

1CFOK One-Compartment, First-Order Kinetics model 

ABA Absolute bioavailability 

aBLM Aquatic Biotic Ligand Model 

ANZECC Australian and New Zealand Environment and Conservation Council 

ARMCANZ Agriculture and Resources Management Council of Australia and New 

Zealand 

ATSDR Agency for Toxic Substances and Disease Registry 

AVS Acid-volatile sulfide 

AWQC Ambient Water Quality Criteria 

AWQS Ambient Water Quality Standards 

BAF Bioaccumulation factor 

BCF Bioconcentration factor 

BF Bioaccessible fraction 

BLM Biotic Ligand Model 

BSAF Biota/sediment accumulation factor 

CBR Critical body residue 

CCA Chromated copper arsenate 

CEC Cation exchange capacity 

CHESS Chemical Equilibria in Soils and Solutions 

CMAQ Community Multiscale Air Quality model 

DEPM Dietary Exposure Potential Model 

DHHS U.S. Department of Health and Human Services 

DL Diffuse Layer model 

DOC Dissolved organic carbon 

DOM Dissolved organic matter 

ECx Effect concentration (for x percent of the study population) 

ECOFRAM Ecological Committee on FIFRA Risk Assessment Methods 

Eh Redox (reduction-oxidation) potential 

EqP Equilibrium partitioning 

FDA Food and Drug Administration 

FIAM Free Ion Activity Model 

FIFRA Federal Insecticide, Fungicide, and Rodenticide Act 

FNB Food and Nutrition Board (NAS) 

GI Gastrointestinal tract 

HFO Hydrous ferric oxide 

HLA Human leukocyte antigens 

HSAB Hard and soft acid and base 

ICP-MS Inductively coupled plasma-mass spectrometry 

IEUBK Integrated Exposure Uptake Biokinetic model 

IRIS Integrated Risk Information System 

IUR Inhalation Risk Unit 

Kd Partition distribution coefficient 

Kow Octanol-water partition coefficient 

LC50 Lethal concentration (for 50 percent of the study population) 

viii 
LIST OF ABBREVIATIONS AND ACRONYMS (continued) 

MAP Metals Action Plan 

MCL Maximum contaminant level 

MPPD Multipathway Particle Dosimetry Model 

NAS National Academy of Sciences 

NASQAN National Stream Quality Accounting Network 

NOM Natural organic matter 

NRC National Research Council 

OM Organic matter 

OSWER Office of Solid Waste and Emergency Response (EPA) 

PBPD Physiologically based pharmacodynamic 

PBPK Physiologically based pharmacokinetic 

PBT Persistent bioaccumulative toxic 

PBTK Physiologically based toxicokinetic 

QSAR Quantitative structure-activity relationship 

RAGS Risk Assessment Guidance for Superfund 

RBA Relative bioavailability 

RDA Recommended dietary allowance 

RDDR Regional Deposited Dose Ratio 

RfC Reference concentration 

RfD Reference dose 

RTC Report to Congress 

SAB Science Advisory Board 

SEM Simultaneously extracted metals 

SHEDS Stochastic Human Exposure and Dose Simulation model 

SPC Science Policy Council (EPA) 

SRWG Science and Research Working Group (SRWG) of the Non-Ferrous 

Metals Consultative Forum on Sustainable Development 

STORET Storage and Retrieval data system 

TCLP Toxicity Characteristic Leaching Procedure 

TMDL Total Maximum Daily Load 

TRIM Total Risk Integrated Methodology 

TRV Toxicity reference value 

TSCA Toxic Substances Control Act 

U.S. EPA U.S. Environmental Protection Agency 

WER Water-effect ratio 

WHAM Windermere Humic Aqueous Model 

WHO World Health Organization 

WHO/IPCS World Health Organization/International Programme on Chemical Safety 

ix 
PREFACE 

Many U.S. Environmental Protection Agency (EPA or the Agency) programs are 

designed to develop guidelines on how to regulate metals. In this process, decisions can range 

from setting environmental release standards, to establishing protective levels in different 

environmental media, to setting priorities for programmatic or voluntary efforts. A fundamental 

input to the decision-making process for most EPA programs is an assessment of the potential 

risks to human health and the environment. 

EPA’s Science Policy Council (SPC) recognizes that metals present unique risk 

assessment issues. The SPC tasked an Agency workgroup, under the auspices of EPA’s Risk 

Assessment Forum, the challenge of devising a Metals Action Plan (MAP) to establish a process 

for ensuring the consistent application of scientific principles to metals risk assessment. The 

MAP included brief descriptions of the Agency’s current activities on metals, identified critical 

scientific issues that need addressing, and recommended the development of a metals risk 

assessment framework. The MAP stated that the Framework should offer general guidance to 

EPA programs for considering the various properties of metals, such as environmental chemistry, 

bioavailability, and bioaccumulation. 

Because of the scientific complexity of metals-specific risk assessment, the Agency 

recognized the need to include stakeholders and the public in the Framework development 

process and to involve experts throughout the Agency. A stepwise process was initiated, 

beginning with the creation of MAP and continuing with framework development and review. 

Workshops and peer-review activities were conducted at regular intervals during development of 

the Framework to ensure that current, accurate science supports the program applications. To 

gain additional information, the Agency contracted for the development of issue papers on 

important topics in metals assessment. These activities, along with input from other federal 

agencies and review by EPA’s Science Advisory Board (SAB), provided additional 

improvements to the Framework drafting process. Additional details on these activities are 

provided below. 

MAP Stakeholder Input. In February 2002, MAP convened a meeting to gather 

stakeholder input to help EPA formulate the plan for developing the Framework for Metals Risk 

Assement. EPA solicited input on organization and content and received comments that were 

adopted into the document with the greatest practical extent. The meeting report and comments 

are available on EPA’s Web site at 

http://cfpub2.epa.gov/ncea/raf/recordisplay.cfm?deid=51737 and 

http://cfpub2.epa.gov/ncea/raf/recordisplay.cfm?deid=51736. 

Science Advisory Board Review. In September 2002, EPA’s SAB reviewed the MAP 

and provided comments. By this point, MAP had some of the panel’s recommendations are 

x 
summarized below; the full text of all recommendations is available at 

http://www.epa.gov/sab/pdf/ecl03001.pdf. 

• The panel agreed that inorganic metals should be assessed differently from organic 

pollutants in a number of contexts. Metals are elements and, although they do not 

degrade, they have complex environmental chemistry. Moreover, some metals are 

essential for living organisms, and metals occur naturally in the environment. 

• The panel agreed that chemical speciation, bioavailability, bioaccumulation, and 

toxicity are key issues in assessing the hazards of metals, with qualifications. 

• The panel recommended consideration of stability and environmental residence times, 

as well as overall environmental chemistry, to determine temporal characteristics of 

metal hazards. 

• The panel recommended greater emphasis on the combined effects of metals, 

including nutritional and toxicological considerations. 

Issue Paper Topics and Science Questions. As part of the development of the 

Framework, issue papers were developed to discuss key scientific topics pertaining to metals. 

The issue paper authors were asked to expand on these topics, with focus on decision-making 

applications, Framework-specific uses, and research needs. The papers, available at 

http://cfpub.epa.gov/ncea/raf/recordisplay.cfm?deid=86119, address the following topics and 

primary questions: 

• Environmental chemistry. How can environmental chemistry be better incorporated 

into assessments for inorganic metals? 

• Bioavailability and bioaccumulation of metals. What methods or tools can be used 

now to reflect metal bioavailability? What scientifically based approaches can be 

used to determine metal bioaccumulation? 

• Metal exposure assessment. What are the relevant exposure pathways for inorganic 

metals to humans and ecological endpoints? 

• Human health effects. What populations are most susceptible to effects from 

inorganic metals? How should toxicity tests be conducted and interpreted, including 

issues of essential elements and dietary salts, among others? 

• Ecological effects. What ecological system characteristics promote increased 

toxicity from metals? 

Peer Consultation Workshop. A draft framework was completed in July 2004, and a 

peer consultation workshop was held from July 27-July 28 to seek input from expert scientists in 

the metals risk assessment field. Workshop participants were from academia; industry, state, 

federal, and Canadian agencies; and various EPA program offices. Stakeholder comments were 

xi 
also received for consideration. Based on comments received at the workshop, the Agency 

contracted with a few workshop participants to explore several issues and insufficiencies 

identified in the human health and environmental chemistry discussions. The Framework draft 

was revised and made available for inter-Agency review. 

Inter-Agency Review. Based on comments received, the framework was revised. 

SAB Review and Public Comment. The SAB review was conducted in February 2005, 

and the final draft of the Framework for Metals Risk Assessment document was completed 

January 25, 2006. The SAB final report is available on the SAB Web site at 

http://www.epa.gov/sab/pdf/metals_sab-06-002.pdf. The SAB found that the framework covers 

the main areas of concern to risk assessors, that some technical corrections and additions were 

needed, and that the document should be restructured and revised to improve the clarity and 

precision of discussions. Based on these recommendations, the framework was revised to focus 

on principles and general metals assessment issues rather than tools. 

xii 
STEERING COMMITTEE AND CONTRIBUTORS1 

CO-CHAIRS 

Anne Fairbrother, U.S. EPA, Office of Research and Development, National Health and 

Environmental Effects Laboratory, Western Ecology Division, Corvallis, OR 

Randall Wentsel, U.S. EPA, Office of Research and Development, Washington, DC 

STEERING COMMITTEE 

Stephen DeVito, U.S. EPA, Office of Environmental Information, Washington, DC 

Alexander McBride, U.S. EPA, Office of Solid Waste and Emergency Response, Washington, 

DC (Retired) 

David Mount, U.S. EPA, Office of Research and Development, National Health and 

Environmental Effects Laboratory, Duluth, MN 

Keith Sappington, U.S. EPA, Office of Research and Development, National Center for 

Environmental Assessment, Washington, DC 

FORUM STAFF 

Gary Bangs, U.S. EPA, Office of the Science Advisor, Risk Assessment Forum, Washington, 

DC 

Marilyn Brower, U.S. EPA, Office of the Science Advisor, Risk Assessment Forum, 

Washington, DC 

Elizabeth Lee Hofmann, U.S. EPA, Office of the Science Advisor, Risk Assessment Forum, 

Washington, DC 

Pamela Noyes, U.S. EPA, Office of the Science Advisor, Risk Assessment Forum, Washington, 

DC 

Seema Schappelle, U.S. EPA, Office of the Science Advisor, Risk Assessment Forum, 

Washington, DC 

William Wood, U.S. EPA, Office of the Science Advisor, Risk Assessment Forum, Washington, 

DC 

CONTRACTOR SUPPORT FOR FINAL DOCUMENT 

Patrick J. Downey, FTN Associates, Inc., Little Rock, AR 

Lynne Haber, Toxicology Excellence for Risk Assessment, Cincinnati, OH 

Ingeborg Harding-Barlow, Palo Alto, CA 

Charles A. Menzie, Menzie-Cura & Associates Inc., Chelmsford, MA 

Kent Thornton, FTN Associates, Inc., Little Rock, AR 

Metals issue paper authors and peer consultation workshop participants are listed in Appendix A. 

1

 Contributors and reviewers are listed in Appendix A. 

xiii 
EXECUTIVE SUMMARY 

The Framework for Metals Risk Assessment is a science-based document that addresses 

the special attributes and behaviors of metals and metal compounds to be considered when 

assessing their human health and ecological risks. The document describes basic principles to be 

considered in assessing risks posed by metals and is intended to foster consistency in how these 

principles are applied across the Agency’s programs and regions when conducting these 

assessments. Although the audience for the Framework is primarily intended to be Agency risk 

assessors, it will also communicate principles and recommendations for metals risk assessment 

to the stakeholders and the public. This guidance will be used in conjunction with guidance 

developed by the programs and regions for use in site-specific risk assessments, criteria 

derivation, ranking or categorization and other similar Agency activities related to metals. 

The purpose of this document is to present key guiding principles based on the unique 

attributes of metals (as differentiated from organic and organometallic compounds) and to 

describe how these metals-specific attributes and principles may then be applied in the context of 

existing EPA risk assessment guidance and practices. While organic compounds, for example, 

undergo bioaccumulation, there are unique properties, issues, and processes within these 

principles that assessors need to consider when evaluating metal compounds. Furthermore, the 

latest scientific data on bioaccumulation do not currently support the use of bioconcentration 

factors and bioaccumulation factors when applied as generic threshold criteria for the hazard 

potential of metals. 

While the science surrounding the metals risk assessment principles continues to be 

studied intensively and evolving rapidly, some areas still lack sufficient information for a 

quantitative assessment to be carried out. Thus, specific approaches may become outdated or 

may otherwise require modification to reflect the best available science and others may be 

addressed only qualitatively until additional information becomes available. Regardless, the 

following principles are more generalized, fundamental properties of metals and should be 

addressed and incorporated into all inorganic metals risk assessments: 

• Metals are naturally occurring constituents in the environment and vary in 

concentrations across geographic regions. 

• All environmental media have naturally occurring mixtures of metals, and metals are 

often introduced into the environment as mixtures. 

• Some metals are essential for maintaining proper health of humans, animals, plants, 

and microorganisms. 

• Metals, unlike organic chemicals, are neither created nor destroyed by biological or 

chemical processes; although, these processes can transform metals from one species 

xiv 
to another (valence states) and can convert them between inorganic and organic 

forms. 

• The absorption, distribution, transformation, and excretion of a metal within an 

organism, depends on the metal, the form of the metal or metal compound, and the 

organism’s ability to regulate and/or store the metal. 

This Framework document is not a prescriptive guide on how any particular type of 

assessment should be conducted within an EPA program office. Rather, it is intended to outline 

key metal principles and how they should be considered in existing human health and ecological 

risk assessment practices to foster consistency across EPA programs and regions. Chapter 1 

identifies the purpose, audience, and scope of the Framework; includes a discussion of typical 

programmatic or regulatory contexts under which the Agency might undertake a metals risk 

assessment; and provides an overview of the metals principles. In particular, the context for risk 

assessment is an important factor in determining the type of analysis appropriate for a particular 

situation. To provide a context for discussion of the framework principles for metals, EPA has 

defined three general categories of assessments: national ranking and categorization, nationallevel assessments, and site-specific assessments. Each type of assessment can vary in level of 

detail from simple, screening analysis to complex, definitive assessments. Approaches and 

methodologies may be appropriate for some or all the risk assessment types and may be more or 

less applicable to screening or complex assessments. Here and elsewhere, the Framework 

acknowledges that data may not be available to implement all the steps in a metals risk 

assessment (e.g., lack of information about metal speciation in some environmental media), 

requiring use of assumptions and a discussion of how such uncertainty influences the risk 

outcome. 

Chapter 2 reviews the human health and ecological risk assessment paradigm and applies 

the metals principles to each phase of the risk assessment process using assessment questions. 

The chapter presents a metals conceptual model showing the interrelationship between the metals 

or metal compounds of interest and the assessment process. Chapter 3 describes how 

environmental chemistry issues affect the assessment of metals. Chapter 4 describes how the 

metals principles should be considered when conducting human health risk assessments, and 

Chapters 5 and 6 describe how the metals principles should be considered in the context of 

aquatic and terrestrial risk assessments, respectively. 

This document discusses scientific issues and scientific approaches. It is intended to 

foster the consistent application of methods and data to metals risk assessment in consideration 

of the unique properties of metals. The Agency will be analyzing the science policy implications 

and developing appropriate policy approaches that are protective of human health and the 

environment. 

xv 
The Framework is the result of contributions from a variety of individuals inside and 

outside the Agency. Their combined expertise and enthusiasm have improved the technical 

quality of the document and its applicability for various risk assessment activities. 

xvi 
1. INTRODUCTION 

Inorganic metals and metal compounds have unique characteristics that should be 

considered when assessing their risks. Some of these characteristics typically are not considered 

when assessing the risks of organic substances. For example, metals are neither created nor 

destroyed by biological or chemical processes; they are transformed from one chemical form to 

another. Native (zero valence) forms of most metals and some metal compounds are not readily 

soluble, and as a result, toxicity tests based on soluble salts may overestimate the bioavailability 

and toxicity of these substances. Some metals (e.g., copper [Cu], selenium [Se], and zinc[Zn]) 

are nutritionally essential elements at low levels but toxic at higher levels, and others (e.g., lead 

[Pb], arsenic [As], and mercury [Hg]) have no known biological functions. Because metals are 

naturally occurring, many organisms have evolved mechanisms to regulate accumulations, 

especially accumulations of essential metals. This metals risk assessment Framework identifies 

metals principles that are fundamental truths (or properties) of metals. The metals principles 

should be addressed and incorporated into inorganic metals risk assessments. 

Metals Principles 

• Metals are naturally occurring constituents in the environment and vary in concentrations 

across geographic regions. 

• All environmental media have naturally occurring mixtures of metals, and metals are often 

introduced into the environment as mixtures. 

• Some metals are essential for maintaining proper health of humans, animals, plants, and 

microorganisms. 

• The environmental chemistry of metals strongly influences their fate and effects on human 

and ecological receptors. 

• The toxicokinetics and toxicodynamics of metals depend on the metal, the form of the metal 

or metal compound, and the organism’s ability to regulate and/or store the metal. 

Because the majority of compounds assessed by the United States Environmental 

Protection Agency (EPA or the Agency) are organic substances, the various guidance documents 

provided for risk assessments of either human health or ecological receptors lack specificity on 

how to account for these and other metal attributes. This document attempts to fill this gap in 

current guidance. 

1.1. PURPOSE AND AUDIENCE 

The primary purpose of the Framework for Metals Risk Assessment is to identify key 

principles that should be addressed in any inorganic metals analysis and to provide EPA program 

1-1
offices and regions with guidance on how to consider these principles in EPA risk assessment 

practices. Although the primary audience will be Agency risk assessors, the Framework will 

also communicate these principles to stakeholders and the public. The Framework relies on the 

draft Framework document, the issue papers, and Science Advisory Board (SAB) comments. 

The issue papers were developed, under EPA commission, to address key scientific topics 

pertaining to inorganic metals. The papers are available on EPA’s Web site at 

http://cfpub.epa.gov/ncea/raf/recordisplay.cfm?deid=86119, which includes links to the draft 

framework (http://cfpub.epa.gov/ncea/raf/recordisplay.cfm?deid=88903). The SAB’s comments 

are available at http://www.epa.gov/sab/pdf/metals_sab-06-002.pdf. 

This Framework document has been developed to supplement previous guidance for use 

in site-specific risk assessments; criteria derivation, ranking, or categorization; and other similar 

Agency activities related to metals. The Framework is not a prescriptive guide on how any 

particular type of assessment should be conducted within an EPA program office. It is, however, 

intended to address issues that are unique to metals and frequently encountered when conducting 

a metals-specific risk assessment. This document does not address issues and methods that are 

common for both metals and organic compounds nor does it develop further guidance on issues 

that remain controversial or unresolved for assessments of risks from chemicals in general. 

Information on general risk assessment topics is available on EPA’s Web site at 

http://cfpub.epa.gov/ncea/ and http://cfpub.epa.gov/ncea/raf/index.cfm. 

The Framework is intended to be used for guidance only. It does not establish any 

substantive “rules” under the Administrative Procedure Act or any other law and will have no 

binding effect on EPA or any regulated entity. Rather, it represents a nonbinding statement of 

policy. EPA believes that the Framework provides a sound, up-to-date presentation of 

principles; provides guidance on how to consider these principles in assessing the risk posed by 

metals; and enhances application of the best available science in Agency risk assessments. 

However, EPA may conduct metals risk assessments using approaches that differ from those 

described in the Framework for many reasons, including, but not limited to, new information, 

new scientific understandings, and new science policy judgments. While the science 

surrounding metals risk assessment continues to be studied intensively and is evolving rapidly, 

some areas still lack sufficient information for a quantitative assessment. Thus, specific 

approaches may become outdated or may otherwise require modification to reflect the best 

available science and others may be addressed only qualitatively until additional information 

becomes available. Application of this Framework in future metals risk assessments will be 

based on EPA decisions that its approaches are suitable and appropriate. These judgments will 

be tested and examined through peer review, and any risk analysis will be modified as deemed 

appropriate. 

1-2
1.2. METALS FRAMEWORK SCOPE 

The Agency regulates metals and their inorganic and organometallic compounds 

(compounds exhibiting properties of both organic and metal compounds) because they have the 

potential to harm human health and the 

environment. The Agency’s SAB has stressed Metals and Metalloids of Primary Interest 

the importance of environmental chemistry Aluminum Manganese 

and its relevance to the assessment of both Antimony Mercury (inorganic) 

inorganic and organometallic compounds. Arsenic 

Barium 

Molybdenum 

Nickel 

However, the complexities of addressing all Beryllium Selenium 

types of metal compounds within a single 

Boron 

Cadmium 

Silver 

Strontium 

document would result in a framework that Chromium 

Cobalt 

Tin 

Thallium 

would be difficult to follow or to apply in Copper Vanadium 

specific cases. Because organometallic Iron

Lead 

Zinc 

compounds exhibit properties common to 

both organic substances and metal 

compounds, the properties of both the organic moieties of these compounds and their 

components would need to be addressed. EPA has already developed frameworks and associated 

guidance documents for assessing properties of organic compounds. Therefore, this document 

addresses only the assessment issues associated with inorganic metal compounds. The 

Framework does discuss natural transformation pathways that form organometallic compounds 

and refers the reader to appropriate Agency documentation or research efforts related to relevant 

risk assessment issues. 

In this Framework, the term “metals” refers to inorganic metals and metalloids that may 

pose a toxic hazard and are currently of primary interest to EPA. However, the principles and 

approaches set forth in the Framework are applicable to all metals. In some instances, metal-bymetal considerations are included, either as examples or as ways to highlight particular 

exceptions. Furthermore, in some cases, this document may discuss particular tools or methods 

that expand on a particular principle and its consideration in the context of EPA hazard and risk 

assessment. The discussions are intended to be illustrative and are not intended to provide a 

complete description of the applications and limitations of any particular tool or method, 

although proper citations to the open scientific literature are included. Nor does the Framework 

provide an exhaustive summary of all the tools and methods available to risk assessors, as this 

type of analysis is beyond the scope of this document. 

1.3. METALS ASSESSMENT CONTEXT 

The context for the risk assessment is a major factor in determining the type of analysis 

appropriate for any particular situation. The Agency conducts a variety of assessments, from 

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site-specific risk assessments to national criteria setting and ranking. To provide a context for 

discussion of the Framework principles for metals, this document has defined three general 

categories of metals assessments: national ranking and categorization, national-level 

assessments, and site-specific assessments. (See Figure 1-1 identifying the three categories of 

assessment in the context of the Agency’s statutory authority.) As shown in the figure, nationalscale and site-specific assessments can vary in level of detail—from simple screening analyses to 

complex definitive assessments. For example, in conducting a national-level screening analysis, 

EPA might undertake a screening-level review of a pesticide or new chemical under the Federal 

Insecticide, Fungicide, and Rodenticide Act (FIFRA) or the Toxic Substances Control Act 

(TSCA), and this may lead to a more definitive assessment if the screen cannot rule out a threat 

to health or the environment. Site-specific screening-level assessments might be sufficient to 

support an environmental impact statement, and a more complex analysis might be necessary as 

a part of a Superfund cleanup action. All three categories of assessments share common 

elements and rely on accurate information and knowledge about how metals behave in the 

environment and when they come into contact with humans or other organisms of concern. 

Metals have unique environmental and toxicological properties that may confound such 

assessments if they are not given consideration. Each of the three general assessment categories 

is discussed in more detail below. 

1.3.1. National Ranking and Categorization 

EPA may rank or categorize some chemicals based on their potential to cause risk to 

human health or the environment. Although there continue to be gaps in data to understanding 

the chemistry, environmental behavior, toxicity, and exposure potential for many chemicals, 

EPA is tasked with protecting and mitigating exposures and harmful effects associated with 

exposure to these chemicals. The Agency often is in a position, despite imperfect and 

incomplete databases, where methods and tools need to be developed to identify, characterize, 

and in some cases, rank and categorize chemicals. 

With more than 80,000 chemicals currently listed on the TSCA inventory that can legally 

be used in commerce within the U.S. (not including pesticides or chemicals that are created as 

byproducts during industrial processes), the Agency needs a way to prioritize substances for 

review or action. Many of the statutes administered by EPA provide specific lists of chemicals 

that require consideration, but often those lists are based on information and analyses previously 

developed by EPA. In addition, the statutes generally provide for adding or deleting chemicals 

from the initial list on the basis of their potential threat to human health or ecological receptors. 

Consequently, a need exists for methods that rapidly screen chemicals for placement on lists or 

that prioritize potentially hazardous substances. 

1-4
Categories of Metals Assessments and Applicable

EPA Statutory Framework

Site-Specific Assessments 

[CWA, CAA, CERCLA, RCRA] 

National-Scale Assessments 

Screening-Level

Analyses

Ranking/ Criteria/ Chemical Reviews Categorization Standards Development [TSCA, FIFRA, [EPCRA, CERCLA, RCRA [CAA, CWA, SDWA, CAA, EPCRA] TSCA, FIFRA, CERCLA, RCRA] SDWA, CAA, CWA] 

Screening-Level

Analyses

Complex Analyses 

Complex Analyses 

Figure 1-1. Categories of metals assessments under EPA statutory 

framework. 

Some of the ranking and categorization methods used by EPA involve identifying certain 

attributes of chemicals that can then be used as indicators of potential human health or ecological 

risk. Example attributes include human and ecological toxicity, production volume, quantities 

released to the environment, persistence in the environment, mobility in the environment as 

indicated by volatility or solubility, and potential to bioaccumulate in the food chain. Other 

methods, which may be less quantitative, rely more on a combination of expert judgment, 

stakeholder input, and availability of information to determine the priority or categorization of 

chemicals for decision making or other action. Examples of programs where EPA identifies or 

categorizes chemicals for priority action based on human health or ecological concerns include 

the following: 

• Selecting chemicals for the Agency’s Toxicity Characteristic regulation (40 CFR 

261.24) that defines hazardous wastes 

• Establishing reporting thresholds for spills of hazardous materials under Superfund 

• Setting priorities for revisions to the Ambient Water Quality Criteria (AWQC) 

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• Listing chemicals under the Toxics Release Inventory 

• Determining priorities for developing drinking water standards 

• Setting priorities for hazardous air pollutant data collection and assessment, and 

• Setting priorities for reviewing existing chemicals under TSCA. 

This list of needs for 

ranking or categorizing chemicals 

is not comprehensive but is 

illustrative of the activities that 

EPA conducts in this regard. In 

addition, the Agency may set 

national standards and guidelines 

for specific chemicals, including 

metals, as described in the next 

section. 

1.3.2. National-Level 

Assessments 

National-level assessments 

may be performed when the 

Agency is setting media standards 

or guidelines for chemicals (e.g., 

Maximum Contaminant Levels 

[MCLs], National Ambient Air 

Quality Standards, AWQC, 

Superfund soil-screening levels) or 

when the Agency is using risk 

assessments to establish controls 

for environmental releases from 

industry or other sources (e.g., 

hazardous waste listings under the 

Resource Conservation and 

Recovery Act, residual risk 

determinations under the Clean Air Act, and pesticide registrations). These assessments can vary 

in level of detail from simple, screening analysis to complex, data-intensive definitive 

assessments. EPA also is charged with establishing controls on environmental releases based on 

the best available treatment technologies (e.g., maximum achievable control technology for air 

Ambient Water Quality Criteria 

EPA’s Office of Water is charged with developing 

Ambient Water Quality Criteria (AWQC) to support the Clean 

Water Act goals of protecting and maintaining the physical, 

chemical, and biological integrity of U.S. waters. Examples 

of chemical-specific criteria include those designed to protect 

human health, aquatic life, and wildlife. Although AWQC are 

typically derived at a national level, there is a long history 

behind the development of methods to accommodate sitespecific differences in metals bioavailability. For example, 

since the 1980s, aquatic life criteria for several cationic 

metals have been expressed as a function of water hardness 

to address the combined effect of certain cations (principally 

calcium and magnesium) on toxicity. Recognizing that water 

hardness adjustments did not account for other important 

ions and ligands that can alter metals bioavailability and 

toxicity, EPA developed the water effect ratio (WER) 

procedure as an empirical approach for making site-specific 

bioavailability adjustments to criteria (U.S. EPA, 1994a). This 

approach relies on comparing toxicity measurements made 

in site water with those made in laboratory water to derive a 

WER. The WER is then used to adjust the national criterion 

to reflect site-specific bioavailability. More recently, the 

Office of Water has been developing a mechanistic-based 

approach for addressing metals bioavailability using the 

Biotic Ligand Model (BLM) (Di Toro et al., 2001; Santore et 

al., 2001; U.S. EPA, 2000b). This model, discussed in 

Chapter 5, predicts acute toxicity to aquatic organisms on the 

basis of physical and chemical factors affecting speciation, 

complexation, and competition of metals for interaction at the 

biotic ligand (i.e., the gill in the case of fish). The BLM has 

been most extensively developed for copper and is being 

incorporated directly into the national copper aquatic life 

criterion. The BLM is also being developed for use with other 

metals, including silver. Conceptually, the BLM has appeal 

because metals criteria could be implemented to account for 

predicted periods of enhanced bioavailability at a site that 

may not be captured by purely empirical methods, such as 

the WER. 

1-6
emission and best available treatment technology for surface water discharges and for hazardous 

wastes). However, even though the standards are based on technological achievability, the 

Agency typically performs risk assessments 

in support of these regulations to help Hazardous Waste Listing Determination 

Under the Resource Conservation and inform management decisions and for use in Recovery Act, EPA is required to make formal 

cost/benefit analyses. decisions on whether to designate certain specific 

industry waste streams as hazardous. For waste Differing environmental conditions streams that are listed as hazardous, the 

across the country affect the generators and handlers of those wastes must 

comply with a comprehensive set of management 

biogeochemistry of metals, making it and treatment standards. In determining whether 

difficult to set single-value national criteria to list a waste as hazardous, the Agency evaluates 

the ways in which that waste is currently being 

that represent the same risk level across the managed or could plausibly be managed by the 

whole country (national standards that apply generators and handlers of the waste. The Agency 

also assesses the physical and chemical 

at the point of exposure, such as MCLs, are composition of the waste. Based on the waste 

characteristics and management practices, EPA less affected by these factors). To conduct then conducts an analysis to determine whether 

such assessments, the Agency commonly potentially harmful constituents in the waste might 

be released and transported to human or undertakes several approaches. One is to ecological receptors. In conducting these 

define one or more exposure scenarios and analyses, the Agency evaluates the potential for 

constituents in the waste material to be released to 

to conduct a relatively detailed analysis. air, surface water, soil, and ground water. It then 

The difficulty in this approach is in selecting models the fate and transport of those constituents 

to potential receptors. 

the appropriate scenario; typically, the 

Agency tries to ensure that the scenario is 

sufficiently conservative to be protective of the population at highest risk (such as populations 

exposed above the 90th percentile) without being so conservative that the standards are 

protective of hypothetical individuals whose calculated risks are above the real risk distribution. 

In selecting the appropriate scenario, the Agency needs to consider all the factors that may affect 

potential risk, including environmental factors affecting the fate, transport, exposure potential, 

and toxicity of the chemicals released. 

Another common approach for a national assessment or criteria derivation is to conduct a 

probabilistic analysis (such as a Monte Carlo analysis), wherein the variability of the key factors 

is described by parameter distributions used as inputs to the probability analysis procedure. The 

result is an integrated distribution of potential risk levels. The difficulties related to conducting 

this kind of analysis are in developing appropriate distributions for each parameter and in 

ensuring that adequate attention is paid to potential correlations among key parameters. These 

correlations often are more complex and difficult to describe for metals than for organic 

compounds. 

1-7
1.3.3. Site-Specific Assessments 

Site-specific assessments are conducted to inform a decision concerning a particular 

location and may also support some national regulatory decisions. They can also vary in detail 

from screening-level to complex, definitive-level analyses. 

• Determining appropriate soil cleanup levels at a Superfund site 

• Establishing water discharge permit conditions to meet AWQS, and 

• Determining the need for emission standards for sources of hazardous air pollutants. 

Establishing Water Discharge Permit Conditions 

The Clean Water Act establishes a two-tier process for setting water discharge permit conditions. 

First, all dischargers must meet the technology-based effluent guidelines limitations requirements. 

Second, if those limitations are not adequate to allow the receiving stream to achieve its designated 

water quality standards (WQS), then more stringent limits are developed to ensure that those 

standards are met. WQS are established by the states and consist of a designated use for the water 

body and a set of criteria for individual chemicals that allow that use to be achieved. EPA has 

published national water quality criteria values for the states to use as guidance in setting their 

standards. 

Once the standards that include the criteria have been established and it has been determined 

that the effluent guidelines alone will not be sufficient to allow those criteria to be met, the state 

prepares a wasteload allocation for all the dischargers to that stream segment, including, where 

appropriate, the nonpoint source discharges. The wasteload allocation generally consists of modeling 

the potential impact on the stream from each discharge of the chemicals of concern and then setting 

the allowable discharges to ensure that the criteria for the chemicals are met. The modeling process 

can be quite complex, potentially taking into account the interactions of the ambient stream conditions 

with the chemicals in the discharge, including dilution, chemical transformations, degradation, settling, 

resuspension, and other processes. For metals, stream characteristics such as pH, organic content, 

suspended solids levels, and numerous other factors can significantly affect how the metal will behave 

and affect aquatic life in the stream segment. Therefore, it is important to understand these 

processes in conducting the wasteload allocation. 

An accurate site-specific assessment for an inorganic metal requires knowledge of the 

form (or forms) of the metal as it enters the environment, the environmental conditions affecting 

the metal (climatological conditions, soil geochemistry, water and sediment chemistry, etc.), the 

existence of plants and/or animals in which the metal might bioaccumulate as well as the uptake 

factors for whatever form(s) the metal may be in, plausible pathways and routes of exposures to 

the human or ecological receptors, and the effect the metal will have on target organisms in 

whatever form in which it reaches that organism and its target organ/system. Although many of 

these same principles also affect the risk potential of organic chemicals, models for predicting 

fate, transport, and toxic properties are generally better defined for organic chemicals than for 

metals. 

1-8
1.4. KEY PRINCIPLES TO CONSIDER IN METALS RISK ASSESSMENT 

One of the purposes of this Framework is to present key principles that differentiate 

inorganic metal compounds from other chemicals. These key principles, defined in subsequent 

subsections, warrant careful consideration when assessing the risks to human health and the 

environment associated with exposures to metals or metal compounds and should be addressed 

and incorporated into metals risk assessments to the extent practicable. For example, it is known 

that certain metal compounds bioaccumulate in human tissues and that this bioaccumulation can 

be related to the metals’ toxicity (SAB, 2006). Contributors to the Metals Action Plan (MAP), 

members of the SAB, and external stakeholders, along with various contributors to and authors 

of this Framework, have discussed these metals principles for consideration in the assessment of 

metals. They are visible throughout this document. In Chapter 2, they are broadly discussed in 

the context of the Agency’s risk assessment process. In Chapters 4 through 6, they are discussed 

in the context of human health, aquatic, or terrestrial risk assessment processes. The remainder 

of this chapter discusses these principles in more detail, focusing on the unique properties of 

metal compounds and why these principles are important for metals risk assessments. 

1.4.1. Metals are Naturally Occurring Constituents in the Environment and Vary in 

Concentrations Across Geographic Regions 

Implications for risk assessment include the following: 

• Humans, other animals, and plants have evolved in the presence of metals and are 

adapted to various levels of metals. Many animals and plants exhibit geographic 

distributions that reflect variable requirements for and/or tolerance to certain metals. 

These regional differences in requirements and tolerances should be kept in mind 

when conducting toxicity tests, evaluating risks, and extrapolating across regions that 

differ naturally in metals levels. 

• As a result of industrialization, current levels of metals may be elevated relative to 

levels occurring naturally. Depending on the purpose of the risk assessment, care 

should be taken to understand and distinguish among naturally occurring levels, 

current background levels (i.e., natural and anthropogenic sources), and contributions 

to current levels from specific activities of concern. 

• Because the diets of humans and other animals are diverse, there may be wide 

variability in the dietary intake of some metals (e.g., in seafood), resulting in both 

temporal variability (e.g., spikes after a seafood meal or with life stage) and 

geographic or cultural variability. 

1-9
1.4.2. All Environmental Media have Naturally Occurring Mixtures of Metals and Metals 

are Often Introduced into the Environment as Mixtures 

Implications for risk assessment include the following: 

• Some metals act additively when they are present together, others act independently 

of each other, and still others are antagonistic or synergistic. Such interactions are 

important aspects of assessing exposure and effects. 

• Interactions among metals within organisms may occur when they compete for 

binding locations on specific enzymes or receptors during the processes of absorption, 

excretion, or sequestration, or at the target site. 

• The presence of and amount of other metals are important when conducting and 

interpreting laboratory tests. 

1.4.3. Some Metals are Essential for Maintaining Proper Health of Humans, Animals, 

Plants, and Microorganisms 

Implications for risk assessment include the following: 

• Adverse nutritional effects can occur if essential metals are not available in sufficient 

amounts. Nutritional deficits can be inherently adverse and can increase the 

vulnerability of humans and other organisms to other stressors, including those 

associated with other metals. 

• Excess amounts of essential metals can result in adverse effects if they overwhelm an 

organism’s homeostatic mechanisms. Such homeostatic controls do not apply at the 

point of contact between the organism and the environmental exposure. 

• Essentiality thus should be viewed as part of the overall dose-response relationship 

for those metals shown to be essential, and the shape of this relationship can vary 

among organisms. For a given population, “reference doses” designed to protect 

from toxicity of excess should not be set below doses identified as essential. 

Essential doses are typically life-stage and gender specific. 

1.4.4. The Environmental Chemistry of Metals Strongly Influences Their Fate and Effects 

on Human and Ecological Receptors 

Unlike organic chemicals, metals are neither created nor destroyed by biological or 

chemical processes. However, these processes can transform metals from one species to another 

(valence states) and can convert them between inorganic and organic forms. Metals also are 

present in various sizes, from small particles to large masses. Implications for risk assessment 

include the following: 

• The form of the metal (chemical species, compound, matrix, and particle size) 

influences the metal’s bioaccessibility, bioavailability, fate, and effects. 

1-10
• The form of the metal is influenced by environmental properties, such as pH, particle 

size, moisture, redox potential, organic matter, cation exchange capacity, and acid 

volatile sulfides. 

• Certain forms of metals are used for evaluating exposure and effects. For example, 

the free metal ion is used for exposure assessments based on competitive binding of 

metal to specific sites of action. 

• Metals attached to small airborne particles are of primary importance for inhalation 

exposures, although a few metals and metal compounds may exist as vapors (e.g., 

mercury). 

• Information developed on the fate and effects of one form of a metal may not be 

directly applicable to other forms. 

• Organometallic forms have different characteristics from inorganic metals and metal 

compounds, and the same general principles and approaches for risk assessment do 

not apply. 

1.4.5. The Toxicokinetics and Toxicodynamics of Metals Depend on the Metal, the Form of 

the Metal or Metal Compound, and the Organism’s Ability to Regulate and/or Store 

the Metal 

These processes are often highly dynamic (e.g., vary according to exposure route and 

concentration, metal, and organism) and thus exert a direct influence on the expression of metal 

toxicity. Implications for risk assessment include the following: 

• Certain metal compounds are known to bioaccumulate in tissues and this 

bioaccumulation can be related to their toxicity. 

• The latest scientific data on bioaccumulation do not currently support the use of 

bioconcentration factor (BCF) and bioaccumulation factor (BAF) values when 

applied as generic threshold criteria for the hazard potential of inorganic metals in 

human and ecological risk assessment (e.g., for classification as a persistent 

bioaccumulative toxic [PBT] chemical). 

• Single value BAF/BCFs hold the most value for site-specific assessments when 

extrapolation across different exposure conditions is minimized. 

• For regional and national assessments, BAF/BCFs should be expressed as a function 

of media chemistry and metal concentration for particular species (or closely related 

organisms). 

• Trophic transfer can be an important route of exposure for metals, although 

biomagnification of inorganic forms of metals in food webs is generally not a concern 

in metals assessments. 

1-11
• Kinetic-based bioaccumulation models (e.g., DYNBAM) have been shown to 

accurately describe bioaccumulation resulting from different exposure routes for 

various metals and aquatic organisms and should be considered as alternatives to the 

BCF/BAF approach when appropriate data are available. 

• Many organisms have developed physiological or anatomical means for regulating 

and/or storing certain metals up to certain exposure levels such that metals may not be 

present in organisms in a concentration, form, or place that can result in a toxic effect. 

• The organ or tissue in which metal toxicity occurs may differ from the organ or 

tissue(s) in which the metal bioaccumulates and may be affected by the metal’s 

kinetics. Target organs may differ by species, mainly owing to differences in 

absorption, distribution, and excretion. Effects at the portal of entry to an organism 

are less dependent on kinetic processes internal to an organism. 

• Both the exposure route and the form of a metal can affect the metal’s carcinogenic 

potential (assessed in the context of human health risk assessment) and its noncancer 

effects. 

• Sensitivity to metals varies with age, sex, pregnancy status, nutritional status, and 

genetics (due to genetic polymorphisms). 

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2. FRAMEWORK FOR METALS RISK ASSESSMENT 

The following discussion addresses issues that are unique to inorganic metals and 

routinely encountered during the inorganic metals risk assessment process. Discussions of issues 

generic to any chemical risk assessments are kept to a minimum because these are dealt with in 

other framework and guidance documents (e.g., U.S. EPA, 2003a, 2000a, 1998a; 

http://www.epa.gov/ncea/ and http://www.epa.gov/ncea/raf). 

This chapter provides an overview of the risk assessment phases and assessment 

questions. Environmental chemistry issues and their implication in the assessment of inorganic 

metals are also discussed. The chapter is organized around the overall risk assessment paradigm. 

(See Figure 2-1, which broadly illustrates the overall risk assessment/risk management process 

and identifies some metals-specific considerations in the problem formulation and analysis 

steps.) An effective risk assessment for metals will account for the unique aspects of metals that 

differentiate them from other substances early and throughout the risk assessment process. 

For assessments of human health or ecological risks at national, regional, or site-specific 

scales, the metals principles can be translated into sets of assessment questions. As appropriate, 

the risk assessor can use these questions to meet the needs of the assessment. The risk assessor 

should consider these questions throughout the risk assessment process; however, they are 

especially important in focusing the assessment during the Planning and Problem Formulation 

phase. Suggested assessment questions are given within this Framework for Problem 

Formulation, Analysis, and Risk Characterization. These questions are not exhaustive but 

provide the risk assessor with a feel for the proper questions to ask. 

2.1. HUMAN HEALTH AND ECOLOGICAL RISK ASSESSMENT: PLANNING 

AND PROBLEM FORMULATION 

Planning and Problem Formulation are critically important for both human health and 

ecological risk assessments (U.S. EPA, 2003a, 2000a, 1998a). The concepts embodied in 

Planning and Problem Formulation are valuable starting points for any risk assessment involving 

metals. Planning and Problem Formulation provide an opportunity for initial consideration of 

the metals’ characteristics and their chemistry. These considerations, along with other aspects of 

the assessment, contribute to the development of a conceptual model that conveys the important 

elements of the metals risk assessment. 

Although Problem Formulation is not explicitly included in the human health risk 

assessment (HHRA) paradigm, as it is in the ecological risk assessment guidelines, current 

practice is to consider many of the issues in the planning stages that the assessor anticipates will 

be incorporated later in the HHRA. This is particularly true for more complex assessments that 

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Figure 2-1. Risk assessment/risk management process for metals. 

2-2
consider multiple metals, pathways/routes of exposure, etc., as is advised in EPA’s Framework 

for Cumulative Risk Assessment (U.S. EPA, 2003e). These planning and scoping activities may 

include 

• defining the geographic scale and scope (site, national-scale, etc.) of the assessment 

• identifying potentially exposed populations and sensitive subpopulations 

• characterizing exposure pathways and exposure routes (conceptual model) 

• describing how exposure will be assessed 

• determining how hazard and the receptor’s dose-response will be assessed, and 

• describing how risks will be characterized. 

For metals, the type of assessment (i.e., 

screening or detailed) and the scale of the assessment 

(i.e., site specific, regional, or national) will determine 

how information on metals can be applied in the 

assessment. Site-specific assessments will involve only 

a single geographical area of concern and, therefore, can 

incorporate locally relevant aspects of environmental 

chemistry, background concentrations, and species 

sensitivities. For regional and national-scale 

assessments, more general assumptions about the form of the metal in the environment, 

deposition pathways, uptake and bioavailability parameters, and sensitive species or 

subpopulations are useful. These general assumptions frequently produce results that are 

conservative in their assumptions in an effort to be protective of sensitive species or locations. 

Regardless, the key principles in metals risk assessment should be considered in all risk 

assessments. 

For metals risk assessment, the risk assessor should consider the following examples of 

questions that should be considered during the planning and scoping of the Problem Formulation 

phases: 

Translating the Metals Assessment 

Principles into Assessment 

Questions 

Translating the metals principles into 

assessment questions should be 

stressed during planning and problem 

formulation. This step helps ensure that 

the principles have been appropriately 

considered. 

Background Concentrations 

• How should background (natural and anthropogenic) levels for metals be 

characterized for the selected spatial scale of the assessment? 

• Is ecoregion-specific information available or is the use of state averages, or 

distributions, compatible with the level of ecological relevance and certainty required 

by the risk analysis? 

• For ecological risk assessments, are acclimation, adaptation, and tolerance data for 

organisms of concern available and are these issues being considered? 

2-3
Mixtures and Interactions 

• How will interactions affecting uptake and systemic effects be considered? 

• Will issues be considered such as reduction of reactivity and increase in mobility by 

organic compounds that form complexes with metals and possible increases in toxic 

effects of organic compounds that form lipophilic complexes with metals? 

• Will interactions with other metals and with organics (e.g., As and polycyclic 

aromatic hydrocarbons) be addressed? 

Essentiality 

• Are nutritional deficits, which can be inherently adverse and can increase the 

vulnerability of humans and other organisms to other stressors, be included in the 

assessment? 

• How will both toxicity and deficiencies of essential metals be characterized? 

Forms of Metals 

• Since environmental chemistry is a primary factor influencing metal speciation and 

subsequent transport, uptake, and toxicity, how will it be included in the risk 

assessment? 

• How will environmental conditions (e.g., pH and oxidation-reduction reactions) be 

addressed to determine metal speciation and mobility? 

Toxicokinetics/Toxicodynamics of Metals 

• What metal-related responses are of most concern in the health risk assessment? 

• Which sensitive subpopulations should be considered for each metal of concern? 

• How will biotic and abiotic factors that influence the bioavailability and 

bioaccumulation of metals be incorporated into the risk assessment? 

• How will environmental factors that affect metal speciation and metabolic capacity of 

biota to regulate internal metal concentrations (homeostatic controls) be accounted for 

when calculating the bioaccumulation potential of metals? 

2.2. METALS CONCEPTUAL MODEL 

The relationships between the sources, exposure, and effects of metals to human and 

ecological receptors are complex and often are specific to a particular site, environmental 

condition, and receptor organism. Because metals are naturally occurring substances that 

undergo extensive biogeochemical cycling (i.e., are not destroyed but change form), transition 

functions between environmental loadings, media concentrations, exposed receptors, and the 

2-4
final organismal or ecosystem responses are affected by natural processes to a much greater 

extent than those that occur with xenobiotic organic contaminants. The assessor should identify 

these transition functions in the conceptual model for all metals assessments. 

The generic conceptual model depicted in Figure 2-2 shows the interrelationship between 

the metals or metal compounds of interest and the health risk assessment process. It is a 

representation of the actual and potential, direct and indirect relationships between stressors in 

the environment and exposed humans (or particular subpopulations) or ecological entities. The 

conceptual model depicts possible pathways from sources of metals and typical ways in which 

risk is assessed (e.g., on the basis of media concentrations, calculated dose, or residues in 

tissues). This model follows the same format as a typical chemical assessment, but it identifies 

areas (primarily in the transition states between environmental compartments) where metalspecific issues require additional consideration. For simplicity, the numerous environmental or 

biological processes that influence the predominant route of exposure or the physical/chemical 

properties of the metal compounds are not depicted in this model, but such processes would be 

used as inputs to models developed for specific assessments. The bidirectional arrows represent 

the fact that the transition functions (models) can be applied in a prospective manner (i.e., in a 

left-to-right direction to determine risks associated with a given load or exposure) or in a 

retrospective manner (in a right-to-left direction to determine the load or exposure associated 

with a predetermined level of risk). The latter is usually done for generating human and 

ecological quality criteria expressed as media concentrations. 

The goals and scope of a health risk assessment, in addition to the availability of data, 

methods, and resources, are among the most important factors that determine the extent to which 

the key principles specific to metals (given in Section 1.4) can be incorporated into an 

assessment. Generally, health risk assessment endpoints are selected during the Problem 

Formulation phase of a risk assessment based on their relevance to risk management goals, 

societal values and laws, known adverse effects of metals, and endpoints of importance to 

stakeholders. Risk assessors will incorporate the metals principles to a lesser extent in 

screening-level assessments than in detailed risk assessments. Site-specific assessments can 

account for more metal-specific processes (particularly environmental chemistry) than can 

national-level assessments that require generalization across multiple ecoregions. Therefore, it is 

recommended that, when appropriate, regional- or national-level ecological risk assessments be 

subdivided into metal-related ecoregions, referred to as metalloregions (McLaughlin and 

Smolders, 2001), such that protection levels, mitigation goals, and ranking results will be 

appropriate for the suite of species naturally present within each type of controlling environment. 

This is directly analogous to the use of ecoregions when establishing water quality criteria 

(Griffith et al., 1999). The Problem Formulation phase of the assessment should clearly identify 

whether a regional approach is being used and, if so, how the metalloregions are defined in terms 

of species composition and environmental controlling factors. 

2-5
Media-based 

Toxicity Models 

Dose-based (M8) Toxicity Models 

diet

soil, seds., air soil, seds., air (M7) fur, feather etc. 

ground water 

soil, seds., air 

surface water

ground water ground water kidney, gill liver, lung, 

surface water surface water 

Metal 

Metal Metal Media-based Bioaccumulation Residue-ba

Fate & Transport Exposure Distribution Exposure Loadings Models Distribution /TK Models Toxicity Mo

Models to Metal in to Media in Media 

Tissues 

(M2) (e.g., mg/kg/d) (M1) (M3) (M4) 

Dietary Exposure 

Models 

(M6) 

2o consumers 

1o consumers 

producers 

Metal 

Bioaccumulation Distribution 

Food Web Model in 

Diet 

(M5) 

Figure 2-2. Generic conceptual model for metals risk assessment. 

2-6
This concept of regional-based ecological assessments is significantly less important in 

human health assessments. In these assessments, the environmental controlling factors (pH, 

water hardness, etc.) may be important determinants in exposure calculations for dietary or 

drinking water exposures. However, to our knowledge, humans have not adapted to particular 

areas of metal enrichment or impoverishment but, rather, choose to live in all environments. 

Therefore, the differences in human sensitivity that should be considered are not geospatially 

correlated. Rather, the assessor should strive to identify potentially sensitive subpopulations, 

such as the very young or the elderly, subpopulations with genetic predispositions to metal 

sensitivity (e.g., Wilson’s disease), or other similar groups. Again, the scope of the Problem 

Formulation phase should clearly address whether the risk results will be applied on a 

population-wide basis, such that protection is afforded to the most sensitive individuals, or 

whether these groups will be given additional scrutiny and separate risk analyses, such that 

results will be applicable only to the general population. 

Figure 2-2 identifies areas in the conceptual model that stand out as metal-specific issues 

as the transitions between environmental loadings, media concentrations, exposure receptors, and 

the final organismal or ecosystem risk. Because metals are naturally-occurring substances with 

which organisms have evolved, it is particularly important to incorporate the natural processes 

that affect metal mobility, speciation, biogeochemical cycling, and sequestration into the health 

risk assessment. These may differ in details or approach, depending on the environment of 

concern (water, land, and air), the final receptor organisms (humans, animals, and plants), and 

the management goal (i.e., whether the management goal is the health of individuals or the 

maintenance of populations or communities). However, the same basic concepts always arise, 

regardless of the assessment context. 

The conceptual model identifies the following issues, indicates the point within the health 

risk assessment process where they occur, and helps direct the remainder of the health risk 

assessment. 

• M1: Fate and transport models. The partitioning and biogeochemical cycling of 

metals into the various environmental media from the loading source depend on the 

physical properties of the initial form of the material and the particular chemistry of 

the receiving environment. Fate and transport models are useful for estimating metal 

speciation, transition kinetics, partitioning, deposition, and potential resuspension 

within the context of environmental levels of the metal and other inorganic 

substances. These can be very detailed for site-specific assessments, or they can 

provide a potential range of processes that might occur over large, regional scales for 

assessments of a more generic nature (e.g., criteria development or ranking schemes). 

Reviews by Paquin et al. (2003), Allen (2002), and EPA (1997a) include up-to-date 

information with regard to the availability of models appropriate for use in evaluating 

fate and transport of metals in aquatic environments (see Chapter 3 on environmental 

chemistry). 

2-7
• M2: Media-based exposure models. Media-specific exposure models are 

mathematical functions used to calculate the exposure of the organism to metals 

directly from abiotic media (i.e., excluding the food web). Estimating the uptake of 

metals from environmental media into biota follows many of the same processes used 

for organic substances, such as understanding dietary preferences, ingestion rates, 

inhalation rates, and movement patterns. Of particular concern with metals health 

risk assessments is accounting for the differing bioavailability of metal species to 

organisms from different environmental media. Exposure to existing environmental 

levels of metals is another issue of considerable importance in this modeling step. 

Exposure assessment issues are considered separately for human health, aquatic, and 

terrestrial receptors (see Chapters 3, 4, and 5, respectively). 

• M3: Bioaccumulation and toxicokinetic (TK) models. Many organic substances 

require metabolic activation to become toxic or, conversely, to be detoxified and 

excreted. Metals do not. Metals may form complexes with proteins or other carrier 

molecules for distribution to target organs or for sequestration and excretion. Their 

bioaccumulation is tissue-specific (e.g., cadmium [Cd] in kidneys). The natural 

occurrence of metals has led to the development of specific mechanisms for uptake, 

metabolism, distribution/storage, and excretion of metals by organisms. These 

processes can impact the use and interpretation of bioaccumulation data and the 

toxicity of bioaccumulated metal. 

• M4: Residue-based toxicity models. If risk to the organism(s) of concern is to be 

based on an estimate of internal dose, then information about the relationship of 

whole-body (or target organ) residue levels to toxic responses should be reviewed by 

the assessor, either from empirical data or physiologically based toxicokinetic 

(PBTK) models. Because of the processes discussed in the previous paragraph, this 

can be particularly challenging for inorganic metals. Metal speciation in the exposure 

matrix can especially influence this relationship because uptake and organ 

distribution kinetics are likely to differ. When available, critical body residues 

(CBRs) can be used to reduce uncertainties in health risk assessments because they 

account for site-specific bioavailability and multiple exposure pathways. However, 

CBRs for metals can vary widely depending on exposure pathway (food vs. water), 

rate of accumulation relative to the detoxification and sequestration processes, and 

form of bioaccumulated metal. Establishment of a valid residue-response relationship 

is critical for successful application of CBRs (see Sections 5.3 and 6.3). 

• M5: Bioaccumulation/food web model. Movement of metals through the food web 

is complicated by factors of bioaccessibility, bioavailability, essentiality, regulation of 

metals (uptake and internal distribution), detoxification, and storage as well as 

accumulation and the natural adaptive capacity of organisms. While the ability to 

quantitatively address all these factors may be limited at present, the assessor should 

at least qualitatively address the potential impacts. Bioaccumulation and trophic 

transfer of metals does occur. However, biomagnification (i.e., increases in 

concentration through multiple levels of the food web) is rare, with the exception of 

certain organometallic compounds, such as methyl mercury, that can biomagnify 

many orders of magnitude in the aquatic food chain. Discussions of methods for 

estimating bioaccumulation in aquatic and terrestrial food webs are found in Sections 

5.2.5.3 and 6.2.5.2, respectively. 

2-8 
• M6: Dietary exposure models. The assessor must carefully consider the 

bioavailability of metals from food items in models that estimate dietary exposure to 

metals. In ecological risk assessments, the wide variation in feeding modes and 

digestive physiology across species limits the ability to make generalizations with 

metals. Generalizations require knowledge of dietary preferences, trophic structure of 

the community, and ingestion and absorption rates. In human health risk 

assessments, the assessor should consider regional, social, and religious dietary 

preferences. Although this node of the conceptual model differs very little from risk 

assessment approaches for organic substances, some metal-specific generalities about 

the relative importance of exposure pathways can be applied to focus (and simplify) 

the process. For example, the highest accumulation of metals in plants generally 

occurs in the roots, and, except for hyperaccumulator species, most plant trophic 

transfer rates can be assumed to be <1. Therefore, direct toxicity to herbivores is less 

likely than for insectivores or from other dietary pathways, and risk to humans from 

most fruits and vegetables (except roots or green, leafy vegetables) is low. On the 

other hand, plants are quite sensitive to some metals and may die before achieving 

levels high enough to be toxic to animals, thereby affecting them indirectly through 

reduction in food availability. A discussion of dietary exposure assessment issues is 

found in Sections 5.2 and 6.2 for aquatic and terrestrial ecological receptors, 

respectively, and in Section 4.2.5.2 for humans. 

• M7: Exposure-based toxicity model. Calculation of an external dose (oral intake, 

gill binding, etc.) for comparison with toxicity thresholds may depend on information 

about relative bioavailability (RBA), speciation of the metal or metal salt, dietary 

preferences and rates, environmental concentrations, essentiality, and metal 

interactions. Toxicity threshold considerations should be based on comparable 

information, such as appropriate metal species in exposure media, similarly 

acclimated or adapted organisms, similar exposure routes, and appropriate 

combinations of essential metals. Chemical equilibrium models such as MINTEQA2 

(Brown and Allison, 1987) may be useful for characterizing the species of metal that 

is present in particular media, making exposure and effect comparisons more 

comparable. This forms the basis of the biotic ligand model (BLM) approach 

(Di Toro et al., 2001; Santore et al., 2001; Paquin et al., 1999) to defining acute 

aquatic toxicity. 

• M8: Media-based toxicity model. This health risk assessment model compares 

environmental concentrations with organism response functions without calculating a 

body burden or internal dose. It is used more frequently for aquatic and soil-dwelling 

organisms, less frequently for wildlife, and very infrequently for human health 

assessments. Consideration of RBA, trophic transfer rates, dietary preferences, 

existing environmental concentrations, and organism adaptations is important for a 

metals assessment. 

• M9: Population, habitat, ecosystem models. Assessors who carry out Ecological 

risk assessments often ask questions related to population growth, habitat change, or 

ecosystem functions in addition to questions related to risks to individual organisms. 

Most of the models and approaches are similar for both metal and organic substances. 

However, metals and other inorganic substances are among the fundamental 

determinants and delimiters of ecoregions (in conjunction with climate, elevation, and 

2-9
day length associated with latitude). Therefore, knowledge of background levels and 

the adaptation of organisms to differing metal levels are essential in developing 

appropriate risk factors for naturally occurring species. 

In summary, the conceptual model lays out a series of working hypotheses about how the 

metal(s) of concern might move through the environment to cause adverse effects in humans or 

ecological systems. These hypotheses are examined through data analyses, models, or other 

predictive tools to determine the probability and magnitude of the occurrence of unwanted 

effects. The approaches used to accomplish this assessement are discussed in general within 

various Agency risk assessment guidance documents. 

2.3. ASSESSMENT PHASE 

The assessment phase of a health risk assessment is the process of estimating exposure 

and understanding the dose-response relationship between biota and the chemical(s) of interest. 

The additional metals-specific factors should be considered during this phase. As with any 

assessment, at the beginning of the Analysis 

phase, the assessor should critically examine the 

Bioaccessibility, Bioavailability, and 

data and models to ensure that they are Bioaccumulation 

appropriate to the level of detail and site-specific, Bioaccessibility refers to the amount of 

regional, or national application of the assessment environmentally available metal that actually 

interacts with the organism’s contact surface (e.g., 

results. Most of the assessment questions in this membrane) and is potentially available for 

chapter are directed toward assisting the assessor absorption (or adsorption if bioactive upon 

contact). Environmentally available metal is the 

with the collection of the appropriate information total amount of metal that is available for physical, 

chemical, and biological modifying influences to address metal-specific issues for conducting (e.g., fate and transport) and is not sequestered in 

either exposure or effects characterizations. an environmental matrix. 

Bioavailability of metals is the extent to which 

bioaccessible metals absorb onto, or into, and 

2.3.1. Bioavailability across biological membranes of organisms, 

expressed as a fraction of the total amount of 

The bioavailability of metals and, metal the organism is proximately exposed to (at 

the sorption surface) during a given time and consequently, the associated risk vary widely under defined conditions. 

according to the physical, chemical, and Bioaccumulation of metals is the net 

accumulation of a metal in the tissue of interest or biological conditions under which an organism is the whole organism that results from all 

exposed. To the extent that available data and environmental exposure media, including air, 

water, solid phases (i.e., soil, sediment), and diet, 

methods allow, the assessor should explicitly and that represents a net mass balance between 

incorporate factors that influence the uptake and elimination of the metal (SAB, 2006). 

Bioconcentration is the net accumulation of 

bioavailability of a metal into the health risk metal in an organism resulting from direct uptake 

assessment. In situations where data or models from water only, such as through gill membranes 

or other external surfaces. 

are insufficient to address bioavailability 

rigorously, the assumptions made regarding 

2-10
bioavailability should be clearly detailed in the health risk assessment as should the associated 

impact on results. 

Although bioavailability may be a defined measurement when considered in certain 

vertebrate animals where metal uptake is directly a function of the concentration of metal in the 

diet, it is not as simple in many other aquatic and terrestrial organisms where food consumption 

is difficult to measure and where metals are present in the surrounding environment and 

available for uptake via nondietary pathways. In this case, as discussed in Meyer (2002), metal 

bioavailability may be more of a conceptual term and not a precisely measured parameter. 

Environmental availability refers to the ability of a metal to interact with other 

environmental matrices and undergo various fate and transport processes. Environmentally 

available metal is not sequestered in an environmental matrix, and it represents the total pool of 

metal in a system that is potentially bioavailable at a particular time and under a particular set of 

environmental conditions (i.e., able to contact or enter into an organism). Environmental 

availability is specific to the existing environmental conditions and is a dynamic property, 

changing with environmental conditions. The bioaccessible fraction of metal is the portion 

(fraction or percentage) of environmentally available metal that actually interacts at the 

organism’s contact surface and is potentially available for absorption or adsorption (if bioactive 

upon contact) by the organism. 

The bioaccessibility, bioavailability, and bioaccumulation properties of inorganic metals 

in soil, sediments, and aquatic systems are interrelated and abiotic (e.g., organic carbon) and 

biotic (e.g., uptake and metabolism). Modifying factors determine the amount of an inorganic 

metal that interacts at biological surfaces (e.g., human digestive system, at the gill, gut, or root 

tip epithelium) and that binds to and is absorbed across these membranes. A major challenge is 

to consistently and accurately measure quantitative differences in bioavailability between 

multiple forms of inorganic metals in the environment. 

The bioavailability issue paper authors (McGeer et al., 2004) provided EPA with some 

practical, standard, and defensible recommendations on concepts, terms, and definitions that can 

serve as a paradigm for studying inorganic metals and their bioavailability. Figure 2-3 presents a 

conceptual framework along with further discussion of metals bioavailability and 

bioaccumulation. 

2.3.2. Exposure Characterization 

Exposure characterization describes potential or actual contact or co-occurrence of 

stressors with receptors (U.S. EPA, 1998a). Metal factors incorporated into this portion of the 

health risk assessment include ecosystem and receptor characteristics that affect the movement of 

metals in the environment including atmospheric deposition, their uptake and accumulation in 

humans and other biota, and distribution into target organs. Specific assessment questions 

include the following: 

2-11
Total Metal Concentration 

Soluble fraction 

Toxicoloical 

accumulation 

Site of 

Toxic 

Action 

Environmental availability 

Internal 

Transport 

and 

Distribution 

Particulate fraction 

Other dissolved 

species 

Nonavailable 

metal 

Soluble species A 

Soluble species B 

Soluble species C

S

p

e

ciatio

n

Uptake 

efficiency 

Dietary intake 

Uptake 

efficiency 

Bioaccumulated Metal 

Uptake 

efficiency

a

d

s

orb

e

d

Bioavailabilit

a

b

s

orb

e

d

Membrane 

uptake 

Detoxification 

and Storage 

Bioaccumulation of metal 

Excretion 

Benign 

accumulation 

Essentiality 

Predation 

Foraging Root tip surface Gut surface Respiratory surface

Cationic 

competition 

Bioaccessibility 

Exposure =====================Î Effects 

Physiological 

membrane 

Bioaccessible Fraction (BF) a: 

Percent soluble metal ion 

concentration relative to total 

metal concentration (measured in 

solution near biomembrane) 

Relative Bioavailability (RBA) b: 

Percent adsorbed or absorbed 

compared to reference material 

(measure of membrane dynamics) 

Absolute Bioavailability (ABA) c: 

Percent of metal mass absorbed 

internally compared to external 

exposure (measures systemic 

uptake/accumulation) 

Figure 2-3. Conceptual diagram for evaluating bioavailability processes and 

bioaccessibility for metals in soil, sediment, or aquatic systems. 

a

BF is most often measured using in vitro methods (e.g., artificial stomach), but it should 

be validated by in vivo methods. 

b

RBA is most often estimated as the relative absorption factor, compared to a reference 

metal salt (usually calculated on the basis of dose and often used for human risk, but it 

can be based on concentrations). 

c

ABA is more difficult to measure and used less in human risk; it is often used in 

ecological risk when estimating bioaccumulation or trophic transfer. 

Source: McGeer et al. (2004). 

2-12 
Background Concentrations 

• What data sources are used to estimate background (natural and anthropogenic) 

concentrations? 

• What are the ranges of background concentrations and how do they vary spatially? 

• What degree of certainty exists in estimates of background concentrations? 

Fate and Transport 

• What environmental transport and air deposition models will be used and what are 

their assumptions, limitations, and uncertainties? 

• How will the environmental chemistry (in air, water, and soils) of metals be 

addressed? 

• What deposition scale (local, regional, or national) is important for the metal and 

receptors being considered? 

• What meteorological factors impacting the fate and transport of metals should be 

considered in the health risk assessment? 

Mixtures and Interactions 

• Is exposure to metal mixtures being incorporated into the exposure assessment? If so, 

how is it being addressed? If not, what is the rationale for not addressing metal 

mixtures? 

• What evidence exists to indicate exposure to the metal(s) of concern is affected by 

metal mixtures in the assessment? 

Essentiality 

• For essential metals, will exposure concentrations exceed the nutritional requirements 

(recommended dietary allowance [RDA])? 

• How do the nutritional requirements vary across species and populations in the 

assessment? 

Forms of Metals 

• What forms (chemical species) of metals are likely to occur at the site(s) of interest? 

• What biogeochemical speciation and transformation processes are relevant for the 

assessment? 

• How might these biogeochemical processes impact exposure assessment for the 

metal(s) of concern? 

2-13
• What transport and fate models are relevant for the environment and metals of 

concern? 

Toxicokinetics/Toxicodynamics of Metals 

• What environmental factors have the most influence on the bioavailability of the 

metals of concern? 

• What methods will be used to address bioavailability in the assessment? 

• How is bioaccumulation being assessed or predicted? 

• To what extent are bioaccumulation predictions being extrapolated across species, 

exposure concentrations, locations, or environmental conditions? 

• What are the key assumptions being used to address bioaccumulation and 

bioavailability and how accurate are these assumptions? 

The objective is to produce a complete picture of how, when, and where exposure occurs 

or has occurred by evaluating sources and releases, the distribution of the stressor in the 

environment, and the extent and pattern of contact or co-occurrence with humans or ecologically 

relevant biota. The metal-specific exposure factors discussed in this framework contribute to the 

exposure characterization, but additional issues that are generally applied to all health risk 

assessments also should be considered (although they are not specifically discussed here). For 

the exposure profile to be useful, it should be comparable with the stressor-response relationship 

generated in the effects characterization. 

2.3.3. Characterization of Effects/Hazard Analysis 

To characterize effects or adverse responses to metals, the risk assessor should describe 

how the effects are elicited, link them to the human populations at greatest risk and/or the 

ecological assessment endpoints, and evaluate how they change with varying exposure levels. It 

is particularly important, especially for inorganic metals, to confirm that the conditions under 

which the exposure occurs are consistent with those of the conceptual model. This will ensure 

that the correct metal species is evaluated for its effects on the populations (including the 

vulnerable subpopulation) or endpoints of concern, or that appropriate models are used for 

extrapolating responses among metal species, biota (laboratory to field, or test species to 

humans), or for varying environments (e.g., metalloregions). Assessment questions regarding 

metal-specific factors for effects analyses or hazard assessments include the following: 

2-14
Background Concentrations 

• What is the relationship between environmental (natural and anthropogenic) 

concentrations and toxicologically relevant metal concentrations? 

• For ecological risk assessments, how are acclimation, adaptation, and tolerance issues 

being addressed in the effects analysis? 

• In human health assessments, have concentrations in locally grown or harvested foods 

been taken into account when estimating elevated metal exposures or estimating 

relative bioavailability of metals in foods, soil, or water? 

• How representative are the toxicity test conditions of the environments being 

assessed? 

Mixtures and Interactions 

• Are toxicological effects of metal mixtures being incorporated in the effects 

assessment? If so, how are they being addressed? If not, what is the rationale for not 

addressing the toxicity of metal mixtures? 

• For particular mixtures of inorganic metals, to what degree are their combined effects 

additive, antagonistic, or synergistic? 

• Is mimicry (competitive interactions among chemically similar metals/metalloids) 

important in the assessment? 

• For site-specific assessments, what evidence exists to indicate the toxicity of the 

metal(s) of concern is affected by the presence of other metals? 

Essentiality 

• For essential metals, are nutritional requirements known (e.g., RDA for humans)? 

• What is the range between concentrations required nutritionally and those associated 

with toxicity reference values (e.g., reference concentration [RfC], Ambient Water 

Quality Criteria [AWQC]) or adverse effect levels used in the risk assessment? 

• Are nutritional deficits that can increase the vulnerability of humans and other 

organisms to other stressors being addressed? 

Forms of Metals 

• Which forms of the metals are most toxicologically relevant? 

• How toxicologically comparable are the forms of metals used in the effects and 

exposure assessment? 

• How might assumptions regarding the toxicity of different metal forms impact the 

effects assessment? How accurate are these assumptions? 

2-15
• How will the atmospheric chemistry, transport, and deposition of metals be addressed 

in the assessment? 

• What meteorological factors impact the fate and transport of metals? 

Toxicokinetics/Toxicodynamics of Metals 

• How does toxicity vary for different metal forms found (or likely to be found) in the 

assessment? 

• Which of these are most important and which are incorporated into the effects 

assessment? 

• How are absorption, distribution, metabolism, and excretion addressed for individual 

metals or mixtures of concern? 

Effects analysis results are summarized in a stressor-response profile. The analysis 

addresses the plausibility that effects may occur or are occurring as a result of exposure to the 

metal(s) of concern, and that linkages between measured effects and assessment endpoints can be 

made (this is especially important for ecological risk assessments). Many of these steps in 

effects assessments are not unique to inorganic metals, and so they are not addressed specifically 

in this Framework. 

Although the prediction of toxicity due to dietary exposure to inorganic metals is 

complicated by wide variation in the bioavailability and toxicity of metals, it is a factor that risk 

assessors should consider in metals assessments. Direct approaches to accomplish this include 

quantifying the bioavailable fraction of bioaccumulated metals in consumers (e.g., analysis of 

tissue fractions such as cytosolic metals) and determining metal speciation in the media of 

concern (water, soil, or air). Comparisons of media values can then be made to toxicity 

reference values using the same metal species. Lacking such information, or for higher tier 

assessments, bioassay methods with field-collected media offer another way to assess 

bioavailability, although other than lead exposure in juvenile swine, such methods have not been 

widely standardized. 

2.4. RISK CHARACTERIZATION 

Risk Characterization is the final phase of the health risk assessment and is the 

culmination of the Planning, Problem Formulation, and Analysis of predicted or observed 

adverse effects. It combines the results of the exposure assessment with information on stressorresponse profiles to estimate the likelihood of effects of specified magnitude(s). The risk 

assessor should describe available lines of evidence and conduct (and report) an uncertainty 

analysis. Conclusions presented in the Risk Characterization should provide clear information to 

risk managers that is useful for decision making. There are no metal-specific methods in the 

2-16
Risk Characterization, other than revisiting the metal factors described above to verify that they 

were accorded proper consideration during the Analysis. However, because metal assessments 

are dependent on specific attributes of environmental chemistry and biological responses related 

to the natural occurrence of metals, it is particularly important that the Risk Characterization 

specify the conditions, locations, and time-frame within which the assessment results are 

applicable. 

For risk assessments conducted for regional or national assessments, criteria 

development, or ranking purposes, it should be acknowledged that results will be based on 

organisms and soil types that result in greatest bioavailability and sensitivity. Care should be 

taken, however, that the organism-environment combinations assessed are, in fact, compatible 

with real-world conditions. Relevant assessment questions include the following: 

Background Concentrations 

• What assumptions are made regarding background (natural and anthropogenic) 

concentrations in characterizing metal risks? 

• How sensitive are the risk assessment results to the presence of background 

concentrations (i.e., are background concentrations a major or minor component of 

the risk estimate)? 

• Have metals with generally high background concentrations (e.g., aluminum (Al) and 

iron (Fe) in soil) been appropriately considered in ecological assessments? 

Mixtures and Interactions 

• How sensitive are the risk assessment results to assumptions regarding exposure and 

effects of metal mixtures? 

• To what extent do the methods and assumptions regarding the exposure and effects of 

metal mixtures introduce intentional or unintentional bias in the Risk 

Characterization? 

Essentiality 

• How do risk assessment results compare to levels required to maintain nutritional 

health? 

• How sensitive are the risk assessment results to methods and assumptions used to 

address essentiality? 

• To what extent do the methods and assumptions regarding the exposure and effects of 

essential metals introduce intentional or unintentional bias in the Risk 

Characterization? 

2-17
Forms of Metals 

• How sensitive are the risk assessment results to methods and assumptions used to 

address the different metal forms? 

• To what extent do the methods and assumptions regarding the exposure and effects of 

metal forms introduce intentional or unintentional bias in the risk assessment? 

Toxicokinetics/Toxicodynamics of Metals 

• How sensitive are the risk assessment results to methods and assumptions regarding 

factors affecting bioavailability and bioaccumulation of the metal? 

• To what extent do the methods and assumptions regarding the factors affecting 

bioavailability and bioaccumulation of metals introduce intentional or unintentional 

bias in the risk assessment? 

2-18
3. ENVIRONMENTAL CHEMISTRY, TRANSPORT AND FATE 

3.1. INTRODUCTION AND TERMINOLOGY 

A general review of factors pertaining to the 

chemistry of metals in sediments, soils, waters, and the 

atmosphere is presented in this chapter in the context of 

risk assessment. Because the behavior of metals defies 

simple generalities, understanding the chemistry of the 

particular metal and the environment of concern is 

necessary. However, the factors that control metal 

chemistry and the environmental characteristics used to 

produce estimates of metal fate and effects can be 

generalized. 

Chemical Species 

The International Union of 

Pure and Applied Chemistry 

(IUPAC) defines chemical 

species as chemical compounds 

that differ in isotopic composition, 

conformation, oxidation or 

electronic state, or that in the 

nature of their complexed or 

covalently bound substituents, 

can be regarded as distinct 

chemical species. 

In environmental chemistry, the phase the species occurs in (gas, liquid, aqueous 

solution, mineral, or adsorbed on an interface between phases) generally is also included in a 

complete definition. In the context of the environmental chemistry of metals, chemists speak of 

a metal species as a “specific form of an element defined as to isotopic composition, electronic 

or oxidation state, complex or molecular structure” and phase (Templeton et al., 2000). 

Metal speciation greatly determines the behavior and toxicity of metals in the 

environment. Speciation refers to the occurrence of a metal in a variety of chemical forms. 

These forms may include free metal ions, metal complexes dissolved in solution and sorbed on 

solid surfaces, and metal species that have been coprecipitated in major metal solids or that occur 

in their own solids. The speciation of a metal affects not only its toxicity but also its 

volatilization, photolysis, sorption, atmospheric deposition, acid/base equilibria, polymerization, 

complexation, electron-transfer reactions, solubility and precipitation equilibria, microbial 

transformations, and diffusivity (Bodek et al., 1988). 

The following sections address the application of hard and soft acid and base (HSAB) 

concepts to metal behavior, including the formation of metal complexes, and the importance of 

pH and oxidation-reduction reactions to metal mobility. These sections also discuss the 

occurrence and interactions of the metals of concern in natural media (including surface and 

ground waters, soils and aquatic sediments, and the atmosphere). Metal sorption behavior, aging 

in soils, metal dissolution and transformation and transfer to plants, and methods of determining 

metal speciation in soils and sediments are important topics considered in these sections. 

3-1
3.1.1. Hard and Soft Acids and Bases: The Stability of Complexes 

Complexes are formed between metals 

(acids) and ligands (bases), both in solution Hard and Soft Acids and Bases 

and at the surfaces of minerals and organisms. Hard acids and hard bases. Complexes 

The toxic reaction of organisms to metals can formed between divalent hard acid cations and 

monovalent or divalent hard bases are ionic and 

be directly related to the nature of the metal relatively weak and are often termed “ion pairs.” 

Complexes formed between Be2+ or trivalent

complexes formed in solution and at the hard acids and hard bases tend to be ionic and 

surface of the organism. relatively strong. 

Soft acids and soft bases. Relatively strong A useful concept that helps to explain covalent bonds are formed in complexes 

the strength of metal complexing and metal between soft and borderline soft acid cations 

and soft bases. Ligand-binding sites on the 

toxicity is that of HSAB, which was external or internal surfaces of organisms are 

introduced by Pearson (1973). In this concept, often soft base and thus bond strongly with soft 

and borderline soft acid cations. 

metal cations are Lewis acids and ligands are 

Lewis bases, with the metal cation and ligand 

in a complex acting as electron acceptor and donor, respectively. “Soft” implies that the species’ 

electron cloud is deformable or polarizable and the electrons are mobile and easily moved. Soft 

species prefer to participate in covalent bonding. Hard species are comparatively rigid and 

nondeformable, have low polarizability, hold their electrons firmly, and prefer to participate in 

ionic bonds in complex formation (Langmuir, 1997). Hard acids form strong, chiefly ionic 

bonds with hard bases, whereas soft acids and 

soft bases form strong, chiefly covalent bonds Ligands 

when they form complexes. In contrast, the Ligands are simply anions or molecules that 

form in coordination complexes with metal ions. bonds formed between hard-soft or soft-hard Depending on whether a ligand shares one, two, 

acids and bases are weak, such that their three, or more electron pairs with metals, it is 

called a mono-, bi-, tri-, or multidentate ligand. 

complexes tend to be rare. Table 3-1 For ligands composed of more than one atom, 

summarizes hard and soft acid and base the atom that directly participates in metal-ligand 

binding is called the donor or ligand. 

relationships for the metals of concern. 

Hard metals (hard acids), which are 

the least toxic, preferentially bind with hard 

bases that contain oxygen, forming weaker bonds with soft nitrogen and sulfur species. The 

strength of binding between hard metals and hard ligands is usually a function of pH. Many of 

the hard metals are macronutrients. Soft metals (acids) bind preferentially with soft S and N 

ligands, forming weaker bonds with hard base species such as hydroxide and sulfate. Soft and 

borderline metals, and Mn2+, which is hard, form bonds of decreasing strength with soft ligands 

such as sulfide, generally in the following order: Pb2+>Cu2+ >Cd2+>Co2+ – Fe2+ >Ni2+ > Zn2+> 

Mn2+. The tendency of metals to bind to soft ligands or to organic substrates (which are 

3-2
Table 3-1. Hard and soft acids (metal cations) and bases (ligands) 

Hard acids Al3+, Ba2+, Be2+, Co3+,, Cr3+, Fe3+, Mn2+, Sr2+, U4+ , 

UO2 

2+, VO2+ 

Borderline acids (between hard and soft) Co2+, Cu2+, Fe2+, Ni2+, Pb2+, Zn2+ 

Soft acids Ag+, Cd2+, Cu+, Hg2+, Hg+, CH3Hg+, Tl3+, Tl+ 

Hard bases F-

, H2O, oxyanions: OH-, SO4 

2-, CO3 2-, HCO3 -

, 

C2O4 

2-, CrO4 2-, MoO4 2- HnPO4 n-3

, HnAsO4 

n-3 

, 

SeO4 

2-, H2VO4 -

, NH3, RNH2, N2H4, ROH, RO­, 

R2O, CH3COO-

, etc. 

Borderline bases (between hard and soft) Cl-

, Br-, NO2 

-

, SO3 

2-, HnAsO3 n-3

, C6H5NH2, 

C5H5N, N3-, N2 

Soft bases I

-

, HS-, S2-, CN-, SCN-, Se2-, S2O3 

2-, -SH, -SCH3, ­

NH2, R-, C2H4, C6H6, RNC, CO, R3P, (RO)3P, 

R3As, R2S, RSH, RS­

“R” refers to an organic molecule. 

Source: Adapted from Langmuir (1997); Huheey et al. (1993). 

usually soft) is greatest for soft and borderline metals (soft acids), followed by the hard metals 

(hard acids), typically in the order Pb2+>Cu2+>Cd2+>Zn2+>Ca2+>Mg2+>> Na+ (Pickering, 1986). 

The tendency of metals to form solid phases, such as sulfides in sediments, is also related 

to their HSAB qualities. For example, extremely insoluble metal sulfides are formed in anoxic 

sediments by soft acid metal cations, such as Hg

2+ (log Ksp = –57.25) or Ag+

 (log Ksp = –49.7), 

whereas borderline hard and hard metal cations such as Mn2+ (log Ksp = –19.25) or Fe2+ (log Ksp 

= –22.39) form slightly more soluble, although still highly insoluble, metal sulfides. Solubility 

products for all sulfides except Ag2S are from Di Toro et al. (1990). The product for silver 

sulfide is from Stumm and Morgan (1970). 

3.1.2. Transformations 

As discussed previously, metals and metalloids can exist in the environment in several 

valence forms. They can also exist as organometallic compounds. Organometallic compounds 

are compounds that have a metal/metalloid-carbon bond. The bonds in organometallic 

compounds are generally covalent and between soft acid metals and soft ligands. 

Metal/metalloid transformation processes, such as metal methylation, occur through interactions 

with other chemicals and biota in the environment. Cycling and distribution of organometallic 

compounds between terrestrial, aquatic, and atmospheric phases may be physically, chemically, 

3-3
or biologically mediated. Table 3-2 lists examples of some commonly occurring, 

environmentally stable organometallic compounds. 

Table 3-2. Examples of organometallic compounds 

Metal/metalloid Organometallic compounds 

AS Methylarsenic acid, dimethyl arsenic acid, trimethyl arsine, 

trimethylarsine oxide 

Pb Tetramethyl/ethyl lead, trimethyl/ethyl lead, dimethyl/ethyl lead 

Hg Methyl mercury, dimethyl mercury 

Se Dimethyl selenide, dimethyl diselenide, seleno-amino acids 

Sn Tributyltin, bis(tributyltin) oxide 

As discussed in this section, organometallic methylation and demethylation rates are 

influenced by both the speciation and bioavailability of the metal, the microbial community, and 

a large number of environmental factors, many of which are interrelated. Sulfide and organic 

matter are important environmental variables that significantly affect methylation; however, their 

effect on methylation/demethylation is poorly understood. The dominant variables differ among 

locations and between seasons, although it is clear that methylation is predominantly a 

biologically mediated process. Methylation/demethylation rates are strongly influenced by 

metal/metalloid speciation and bioavailability. 

Environmental methyl-metal concentrations reflect the net methylation rather than simple 

rates of methyl-metal synthesis. Table 3-3 presents metals that are involved in abiotic or biotic 

methylation/demethylation processes. With the exception of As and Se, the metals in the table 

form stable complexes with either methyl or ethyl groups. In addition to methyl/ethyl 

compounds, metals can also be incorporated biochemically into stable organometallic 

compounds (e.g., As into lipids and Se into amino acids). 

Organometallic environmental transformations may affect both the mobility and toxicity 

of metals. The transformation rates and the organometallic products are dependent on 

environmental conditions and the population of microorganisms available. For example, 

methylation/demethylation rates are dependent on the speciation of the metal, the microbial 

community, the environmental variables (e.g., pH, temperature, reduction oxidation potential, 

organic matter, dissolved oxygen, nutrient availability, salinity, complexing agents), and the 

distribution of the metal between compartments (sediment, water, gaseous). The interrelatedness of these processes has made research into unraveling the factors controlling net 

methylation difficult and, to date, incomplete. However, some general trends can be predicted 

with some certainty, as discussed in this section. 

3-4
Table 3-3. Metals/metalloids involved in methylation processes 

Process Metals affected 

Environmentally stable organometallics Si, Ge, Sn, Pb, Hg, As, Sb, Se 

Abiotic chemical methylation Hg, Pb, Sn 

Abiotic demethylation Sn, Pb 

Biotic methylation As, Cd, Hg, Pb, Se, Sn 

Biotic demethylation As, Hg, Sn, Pb 

Source: Bodek et al. (1988). 

3.1.2.1. Biotic Methylation Transformations 

Biotic methylation occurs when organisms, primarily microorganisms, transfer alkyl 

groups to bioavailable metals. In general, it is thought that anaerobic sulfate-reducing bacteria 

are the principal methylators in freshwater and estuarine environments. However, methylation 

rates are not always correlated with sulfate-reducing bacteria. Not all sulfate-reducing bacteria 

are capable of methylating, and the efficiency of methylation is dependent on the activity and 

structure of the bacterial community. Other bacteria may be involved in methylation. Biotic 

methylation occurs predominantly in the sediment column; however, because the water column 

by volume is much larger, water column methylation is important. 

Maximum methylation rates typically occur at the redox boundary, which varies 

seasonally and frequently coincides with the sediment-water interface (Ullrich et al., 2001). 

Methylation rates decrease with increasing sediment depth, probably due to a decrease in biotic 

habitat. Microorganisms may also demethylate (or dealkylate) organometallic compounds. 

Microbial-mediated transformations are frequently the most important environmental 

organometallic processes. Generally, as the amount of organic material increases in a system the 

microbial populations also increase. 

High temperatures and anaerobic conditions generally favor metal-methylation 

formation, and demethylation processes are generally favored under low temperatures and/or 

aerobic conditions. Studies on the effects of pH are not consistent. Interconnected parameters 

include pH effects on the microbial communities and effects on the speciation distribution of the 

metals/metalloids in the water and the sediment as well as adsorption rates. Organometallic 

compounds appear to increase in the water column in low pH environments, but this may be due 

to release of methylated metals from the soil and subsequent depletion of organometallic 

compounds in the soil. Therefore, pH effects on net methylation in a system are not fully 

understood. In freshwater ecosystems, where sulfate concentrations are typically low, increase 

3-5
in sulfate concentration increases methylation rates. However, in reducing environments, 

increasing sulfide concentration decreases methylation rates. General trends in 

methylation/demethylation rates are outlined in Table 3-4. 

Table 3-4. General trends of environmental factors affecting rates of 

methylation/demethylation 

Organometallic 

transformations 

Temperature pH SO4 

-2 Organic 

matter 

Redox Salinity 

High Low High Low High Oxic Anoxic High 

Net methylation ↑ ↓ ? ? ? ? ↓ ↑ ↓

Methylation aq ↑ ↓ ↓? ↑? ↓ ↓↑ ↓ ↑ ↓

Methylation sed ↑ ? ↑ ↓ ↓ ? ? ? ? 

Demethylation ↓ ↑ ↓ ↑ ? ? ↑ ↓ ? 

↑ indicates an increase in rate. 

↓ indicates a decrease in rate. 

? indicates conflicting data or insufficient data to indicate a likely trend. 

The inhibitory effect of sulfide is probably not due to metal sulfide formation but, rather, 

to the formation of less bioavailable metal-sulfur complexes. High organic matter may increase 

abiotic methylation through humic/fulvic metal reactions; however, this mechanism is poorly 

understood and confounded because biotic methylation rates may increase in environments with 

high organic matter. In ecosystems with high dissolved organic carbon (DOC) concentrations, 

DOC may bind with metals/metalloids, rendering them unavailable and thereby reducing biotic 

methylation rates. 

3.1.2.2. Abiotic Transformations 

Some organometallic compounds, those with electronegativities greater than 1.7, are 

stable under environmental conditions. In a few instances, metals can undergo abiotic 

methylation processes. Carbon-metal bonds with more polar (metal electronegatives <1.7) will 

undergo hydrolysis (reaction with water). Abiotic chemical methylation can occur by three 

mechanisms: transmethylation reactions between mercury and tin/lead alkyls, reactions with 

humic/fulvic substances, and photochemical reactions. 

3-6
3.1.3. Aquatic Chemistry 

Exposure and risks associated with metals in aquatic environments depend on the forms 

of the metals and on the factors that influence the presence of these forms as well as on the fate 

and transport of the metals. 

3.1.3.1. Processes 

A number of factors influence the sorptio

of metals in aquatic systems. Speciation/ 

complexation is the distribution of a given 

constituent among its possible chemical forms, 

including metal complexes, which have differing

tendencies to be adsorbed or desorbed; precipitati

is the process by which dissolved species exceed 

the solubility limits of their solids, so that some o

the species precipitate from solution; colloid 

formation can result in metals being sorbed or 

coprecipitated with colloidal-sized particles; biofixation occurs when biological processes 

(usually involving microorganisms or plants) result in the binding of metals to solid materials; 

interactions with natural organic matter can also result in sorption. In addition to these factors, 

sorption is influenced by changes in pH, oxidation potential, salinity, concentrations of 

competing ions, the nature of sorbent phases and their surface areas, and surface site densities. 

n 

Factors Affecting Sorption of Metals 

Speciation/complexation 

Precipitation 

Colloid formation 

Biofixation

Interactions with natural organic matter 

on Changes in pH 

Oxidation potential 

Salinity 

f Competing ions 

Nature of sorbent phases 

Surface site densities 

3.1.3.2. Speciation and Complexes 

Metal species dissolved in water may occur as free ions, or aquo-ions, or as complexes. 

Free metal cations are generally surrounded by coordinating water molecules and so have been 

termed “aquocations,” although by convention the water molecules are ignored when writing 

chemical reactions involving metal cations. 

The total analytical concentration of a given metal in water is the sum of the 

concentrations of its free ion and its complexes and any metal associated with suspended solids, 

whether organic or mineral. For example, the total molal concentration of lead, ΣPb, in a natural 

water might equal: 

ΣPb = mPb2+ + mPbOH+

 + mPbCO3

o

 + mPbHCO3

+

 + mPbSO4

o

 + mPb(suspended solids) 

In most natural waters, the concentration of free lead ion, mPb2+, is less than the sum of 

the concentrations of its complexes, which in this case are lead complexes with hydroxyl, 

carbonate, bicarbonate, and sulfate ions. Other metals that are found in natural waters most often 

3-7
as complexes and not as free ions include Al3+, Ag+

, Cu2+, Fe3+, and Hg2+. The metalloids As 

and Se and the metals Cr, Mo, Sb, and V occur most often in aerobic waters and soils as 

covalently bonded to oxygen in oxyanions. Under oxidizing conditions, these include arsenate, 

selenate, chromate, molybdate, and vanadate, which themselves are complexes. 

Complexes that incorporate metals play a major role in controlling the availability and 

fate of metals in the environment. Increasing the fraction of a metal that is complexed increases 

the solubility of minerals of that metal (Langmuir, 1997). For example, the solubility of lead 

sulfate is related to the molal concentrations of free lead and sulfate ions through the expression 

below: 

Ksp = (γPb.mPb2+)(γSO4.mSO4

2-) 

where the terms γPb and γSO4 are the activity coefficients of the ions. The product of the ion 

activity coefficient and the molal concentration of each species equals the activity of the ion. 

This equation shows that for a given total Pb concentration, the greater the amount of Pb that is 

complexed, the lower the concentration of free Pb ion. This means that as the extent of Pb 

complexing increases, the total Pb concentration must also increase to reach saturation 

equilibrium with lead sulfate. In other words, metal complexing increases total metal solubility. 

Metal complexing also has a direct influence on metal adsorption to organic matter or 

mineral surfaces. For example, metal carbonate, sulfate, and fluoride complexes are usually 

poorly adsorbed, whereas metal hydroxide complexes are strongly adsorbed (Langmuir, 1997). 

In summary, metal complexing generally increases the solubility and mobility of metals in 

surface and ground waters (with the exception of metal hydroxide complexes). 

For many metals, the free metal ion is thought to be the primary metal species that causes 

toxicity to aquatic organisms. This is consistent with the free ion activity model (FIAM), which 

carries the assumption that the free or aquo-ion is the most biologically active form of the 

dissolved metal. If this is true, the key parameters that can modify the degree of toxicity are 

those that affect speciation, such as pH and the amount of inorganic and organic ligands (e.g., 

DOC) that can form metal complexes and so provide alternative binding sites for the metal ion. 

Metal toxicity also will be affected by other dissolved ions (e.g., Na, Ca) that compete with 

metals for binding sites on the gills of fish or on the respiratory surfaces of other aquatic 

organisms. However, it should be noted that many metal complexes (whether bound to organic 

or inorganic ligands) are reversible, particularly if environmental conditions change (e.g., pH 

decreases). 

3-8
3.1.3.3. Importance of pH and Redox Conditions 

The pH is probably the single most important variable that influences the behavior of 

metals in the environment. Thus, metal complexes with sulfate, fluoride, chloride, and 

phosphate are most stable and important below pH 7, whereas metal carbonate and hydroxide 

complexes become increasingly more important above pH 6–8. 

Also, as discussed in Section 3.1.6, hydrogen ion competes with metal cations for 

adsorption sites, so that adsorption of metal cations by hydrous ferric oxide (HFO), for example, 

is low in acid systems but increases with increasing pH. In contrast, oxyanions of As, Mo, Se, 

and Cr tend to be desorbed from HFO with increasing pH because of competition between the 

oxyanions and OH- ion for sorption sites. Furthermore, the solubility of most metal-containing 

minerals is greatest under acid conditions, decreasing with increasing pH. Figure 3-1 shows the 

locus of measured values of oxidation potential (Eh) and pH in aquatic systems. The principal 

controls on Eh are atmospheric oxygen and organic matter. High Eh (oxidizing or aerobic) 

conditions are maintained in the atmosphere and in most surface waters and shallow soils in 

contact with atmospheric oxygen. The lowest Eh values and reducing or anaerobic conditions 

are found in water-logged soils and sediments that contain organic matter and in ground waters 

that contain a few milligrams per liter or more of DOC. Intermediate Eh conditions are found in 

waters and sediments that are only partially oxidized because of their relative isolation from the 

atmosphere. Measured Eh values may not agree with Eh values computed from the 

concentrations of redox-sensitive species. The difference between measured and computed Eh 

values is discussed at length by Stumm and Morgan (1996) and Langmuir (1997). 

3.1.4. Ground Water and Metals Mobility 

Site-specific nonorganic metal health risk assessments for EPA programs often are used 

to predict the rate of movement of metals through soils and their subsequent movement and 

concentrations in ground water. The primary processes governing the environmental fate and 

transport of metals in the subsurface are advection, dispersion, matrix diffusion, and retardation 

(U.S. EPA, 1994b). Advection and dispersion are functions of the system rather than of the 

contaminant. Matrix diffusion, which is a function of the contaminant, is relatively unimportant 

and is omitted in most model transport algorithms. Retardation depends on a number of factors 

(Langmuir, 1997; U.S. EPA, 1994b) and may involve or be affected by the following: 

• Sorption - The attachment of metal species to mineral surfaces or other surfaces. 

• Speciation - The distribution of a given constituent among its possible chemical 

forms, including metal complexes, which have differing tendencies to be adsorbed or 

desorbed. 

3-9
Figure 3-1. Approximate positions of some natural environments in terms of 

Eh and pH. The dashed line represents the limits of measurements in natural 

environments. 

Source: Baas-Becking et al. (1960). 

• Precipitation - The process by which dissolved species exceed the solubility limits of 

their solids, so that some of the species precipitate from solution. When a metal 

species reaches mineral saturation, addition of further amounts of the species to 

solution are precipitated, not adsorbed. 

• Colloid formation - The process of forming colloids and the association of metal 

species with them. The metals may be sorbed or coprecipitated with colloidal-sized 

particles. 

• Biofixation - The binding of metals to solid materials due to the interactions of 

microorganisms or plants. 

• Natural organic matter - (NOM) interactions. 

3-10
• Other processes - Other processes include changes in pH, oxidation potential, 

salinity, concentrations of competing ions, the nature of sorbent phases and their 

surface areas, and surface site densities. 

3.1.4.1. Application of Partition Coefficients to Metal Mobility in Ground Water 

Owing to the complexity and multiplicity of the processes involved, recourse is often 

made to the use of a single partition or distribution coefficient that describes the degree to which 

the contaminant’s transport is retarded relative to water. This approach starts with defining the 

retardation factor: 

ν 

R f ν 

p

= 

c 

where: 

Rf = the retardation factor, 

νp i= the velocity of water through a control volume, and 

νc = the velocity of contaminant through a control volume. 

Langmuir (1997) noted that the retardation factor is related to the distribution coefficient 

through the below expression: 

Rf = 1+ ρb Kd

ne 

where: 

ρb is the porous media bulk density, and 

ne is the effective porosity at saturation given as a volume fraction. 

Laboratory adsorption studies often find that, in simple systems, the value of log Kd for 

metal adsorption increases linearly with pH. For example, for Zn2+ adsorption by HFO, 

Langmuir (1997) noted that adsorption followed the log equation: 

log Kd = !5.48 + 1.77 pH 

The properties and applications of partition or distribution coefficients for metals are 

comprehensively reviewed in EPA (1999a). These coefficients may be obtained from the 

literature, estimated using mathematical models, or measured. Properly designed column 

3-11
experiments generally give results more representative of the in situ behavior of soils and 

sediments than do batch tests (Langmuir, 1997). 

Models using partition distribution coefficients (Kd) have significant inaccuracies for 

metals, and the application of single partition coefficient values for individual metals should be 

limited to site-specific assessments (when they can be calibrated for a specific site); to regionaland national-scale studies where bounds of potential Kd values are adequate; or reasonably 

representative single values that have sufficient functionality built in to account for variability 

due to pH and soil composition. 

EPA (1999a) discusses the advantages and disadvantages of several methods for 

measuring partition coefficients, including laboratory batch testing, in situ field batch testing, 

flow-through testing, and field modeling. In many national assessments, EPA has used the 

MINTEQ model and its subsequent versions to generate generic partition coefficients that may 

be applied to regional or national mobility evaluations (http://www.epa.gov/ceampubl/mmedia/ 

minteq/index.htm or http://www.lwr.kth.se/english/OurSoftware/Vminteq/). 

For screening assessments: 

• Partition coefficients have been tabulated as a function of pH by EPA (U.S. EPA, 

1998b), and the Agency has also presented non-pH-dependent values for Pb (900) (Pb 

values have large variability and selecting meaningful values is difficult), mercuric 

chloride (58,000), and Hg (1,000) (U.S. EPA, 1999b). 

• In simple systems, the value of log Kd for metal cation adsorption usually increases 

linearly with pH, whereas the value of log Kd generally decreases with pH for anion 

adsorption. 

For definitive assessments: 

• It may be possible to estimate metal adsorption with some accuracy without having to 

measure it in the field, depending on the information available on a specific soil, 

surface-water, or ground-water system. To do so, at minimum, the risk assessor will 

need to determine the amounts and surface areas of the potentially sorbing materials 

(e.g., metal oxides, clays, and organic matter [OM]) in a soil or sediment or in 

suspension in a stream, and the detailed chemical composition of the water, especially 

its pH and metal concentration. Literature information can then be used to estimate 

the sorption properties of these materials for use in the diffuse layer (DL) sorption 

model. 

• If greater accuracy or site specificity is required, it may be necessary to measure 

metal adsorption in laboratory experiments designed to parameterize the DL model 

for application to a specific study area. The risk assessor could run experiments 

(batch tests) that attempt to reproduce the composition of waters and sorbing solids in 

the study area. 

3-12
3.1.5. Sediment Chemistry 

In addition to the challenges posed 

by metal chemistry, the sedimentary 

environment is complex and often highly 

heterogeneous. Fortunately, generalizations 

can be made about the sedimentary 

environment in order to progress toward a 

method for risk assessment. This brief 

review summarizes information on the 

composition of sediments; processes that act 

on sediments and their metal burden; and the 

chemistry of the sedimentary environment 

that influences the fate, bioavailability, and 

effects of metals. It is important to consider these factors in light of the aim of estimating 

potential biological effects of metals in sediments. 

Sediment solids can hold up to a million times more metal than an equivalent volume of 

water. The exact proportion of a chemical held by sediment relative to water is a function of a 

metal’s chemistry as well as the chemistry of the sediment solid and the surrounding 

environment. Further, this distribution is dynamic (Diamond and Mudroch, 1990). Because of 

their large capacity to “hold” metals, sediments have been characterized as “sinks.” Sediments 

can serve as temporary sinks from which some of the metal can enter ecological and human 

food webs through several routes (e.g., Diamond, 1995), primarily through accumulation in 

benthic organisms. These organisms include those that fully or partially live in the sediments 

(e.g., tubificids, chironomids, trichopteran larvae) or those that feed from the sediment bed (e.g., 

suckers, carp). Some organisms obtain their chemical dose from both pelagic and benthic routes 

(e.g., lake whitefish, walleye), but because of high chemical concentrations in sediments, the 

benthic route can be the dominant route of uptake (Morrison et al., 2000). 

For humans, the route of entry of metals from the sediments is through water used for 

drinking, bathing, and swimming. The availability of these metals is mediated by sedimentwater exchange processes that can result in the release or remobilization of chemicals from the 

sediment bed. However, owing to the ability of Hg to bioaccumulate in its monomethyl form, 

fish consumption is the critical route of exposure to this metal for humans.

2 

Many important chemical reactions involving the metals of concern occur in the finegrained materials that accumulate in the deep parts of water bodies. The controlling factors or 

master variables that influence metal chemistry are redox potential and pH. A depth profile of 

Sediments 

Bed or bottom sediments are found at the 

bottom of lakes, rivers, and estuaries. 

Sediments have several sources that influence 

their composition and chemistry. The type and 

chemistry of sediments is also determined by 

their location in the water body as well as the 

characteristics of the water body. At any given 

site, metals can be associated with solid-phase 

minerals, organic matter, colloids, and pore 

water. The solid phase can vary from sand (>63 

μm) to silt (2–63 μm) to clay (<2 μm). Because 

clays have more active binding sites than do the 

other grain sizes and because of their high 

surface area-to-volume ratio, clays –or finegrained particles--are of greatest significance in 

terms of metal binding. 

2

 Marine biota can also be a significant route of exposure to arsenic in its organic forms, such as methylarsonic acid; 

however, these arsenic species are significantly less toxic than the inorganic forms (Fowler, 1983). 

3-13
the sediments will reveal decreasing sediment porosity and concentrations of dissolved oxygen 

because oxygen is consumed as organic matter decomposes. The pH is often relatively constant 

or may decrease with depth, but alkalinity may increase owing to mineralization of organic 

matter (Stumm and Morgan, 1996). As dissolved oxygen is consumed, anaerobic microbes use 

other electron acceptors in redox or oxidation-reduction reactions in the order of nitrate, ferric 

iron, ammonium, sulfate, and bicarbonate to produce carbon dioxide, ammonia, sulfide, and 

methane. 

The risk assessor can ascertain the redox status of the sediments by measuring the 

concentration of dissolved oxygen or other redox-sensitive species, by calculating pE (a measure 

of electron availability in solution), or by measuring Eh (millivolts) or the electromotive force of 

the pore water solution. The zone of transition from oxic to anoxic conditions is the redoxycline, 

which can migrate vertically, depending on the mixing of the overlying water column (e.g., 

Diamond and Mudroch, 1990). For example, the redoxycline may be 5-10 cm below the 

sediment-water interface in a well-oxygenated oligotrophic lake or river, but it may be above the 

sediment-water interface in a thermally stratified eutrophic lake or river. 

3.1.5.1. Metal Chemistry in Sediments 

In this discussion, the risk assessor should consider two pools of metals. The first pool 

consists of metals that exist as aqueous (or dissolved) species bound to colloids or DOM and 

those bound to sediment particles through an exchangeable binding process. This pool is often 

referred to as the “exchangeable” or “labile” pool. The second pool consists of metals found 

within the mineral matrix of the sediment solids. This pool is largely unavailable to biota, and its 

release will occur over geologic time scales, through diagenetic processes. Because the latter 

pool is largely unavailable, only the exchangeable pool of metals is considered. Note that the 

exchangeable pool will be composed of naturally occurring metals that are released into solution 

as a result of weathering and diagenetic processes as well as metals released into the 

environment as a result of anthropogenic activities. 

The exchangeable pool of metals is subject to speciation in the aqueous phase (e.g., 

within the pore water) and sorption to solid phases, where sorption is a general term that includes 

adsorption (the accumulation of matter at the solid-water interface or a two-dimensional process) 

and absorption (inclusion in a three-dimensional matrix) (Stumm and Morgan, 1996). Here, 

speciation refers to the distribution of metal species in a particular sample or matrix or species 

distribution (Templeton et al., 2000). In the aqueous phase, metal will react or bind with 

dissolved ligands according to the pH, Eh, ionic strength, and abundance of ligands (see Section 

3.1.3, Aquatic Chemistry). 

The concentration of metal in the dissolved phase is controlled by sorption to the solid 

phase. Although sorption can occur rapidly, desorption or dissolution of metal from the solid 

phase may be a two-phase process, where the second phase is rate limiting (e.g., Nyffeler, 1986; 

3-14
Santschi et al., 1986). If the kinetic limitation of reactions are not considered, the distribution of 

metals among aqueous species and between the aqueous phase and the solid phase can be 

estimated by using one of the speciation models. Several speciation/complexation models are 

available to perform this calculation, such as MINEQL+ (Schecher and McAvoy, 2001), the 

Windermere Humic Aqueous Model (WHAM) (Lofts and Tipping, 1998), and MINTEQA2 

(Allison et al., 1991). These models work well under oxic conditions, but estimates of metal 

binding are less reliable under anoxic conditions, where metal concentrations are most often 

controlled by the solubility of metal sulfides. Furthermore, in some circumstances equilibrium 

may not be achieved, particularly when the redoxycline moves more quickly than the rate of 

metal reaction or when the reaction is governed by microbial processes, as occurs with the 

methylation of Hg or As. 

In oxic sediment pore waters (above the redoxycline), metals will exist as aqueous 

species, that is, as freely dissolved ions or metal complexes (e.g., phosphate, sulfate, or carbonate 

complexes), and associated with colloids. Solid-phase reactions are controlled by iron 

oxyhydroxides and manganese oxides that may exist as colloids, sediment particles or surface 

coatings of particles, OM that may also exist as colloids or coat sediment particles, and clay 

colloids and particles. 

As Eh declines, the solid-phase manganese oxides are the first to be reduced and thereby 

dissolve, which releases metals that have been sorbed or coprecipitated with these minerals. 

Some of the metals released into the pore water may then be adsorbed by iron(III) 

oxyhydroxides, which are the next to dissolve as the Eh continues to drop. Under reducing 

conditions, particularly as sulfate is consumed and the S is converted to sulfide, metal 

concentrations in pore waters again drop as solid-phase metal-sulfides are formed (see discussion 

below about the role of acid-volatile sulfides [AVSs] in regulating toxicity). 

As a result of redox chemistry, metals can undergo seasonal redox-driven cycling 

between the water column and sediments or within the sediments, depending on the position of 

the redoxycline. The stages in the cycling are, first, the adsorption or coprecipitation of metals 

with iron and manganese hydroxides under oxidizing conditions; then with the development of 

moderately reducing conditions, the reduction and dissolution of the manganese and iron 

oxyhydroxides and consequent release of the associated metals into the water or pore water; 

followed by their diffusion upward toward the zone of low metal concentrations under oxidizing 

conditions. It is also possible for dissolved metals to diffuse downward toward the zone of low 

metal concentration owing to their precipitation as sulfides. As a result of this vertical cycling, 

the depth profile of metals in pore water may not match that of the solid phase (e.g., Carignan 

and Tessier, 1985). Moreover, it is possible, but less usual, that the cycling can occur relatively 

rapidly and involve a significant portion of the solid-phase metal. Under these conditions, the 

solid-phase sediment profile reflects this reworking rather than the historical record of metal 

loadings (MacDonald et al., 2000). 

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The pH controls metal speciation and binding by affecting the species distribution of 

dissolved ligands (e.g., phosphate, sulfate, carbonate, humic substances) and the surface charge 

of binding sites on DOM and solid phases such as iron oxyhydroxides. Generally, at low pH, 

when surface sites are protonated, the sorption of cationic metals decreases, and, hence, metal 

mobility increases. The converse occurs at high pH, which results in low metal solubility and 

greater sorption. The patterns of dissolution and sorption are reversed for metalloids, such as As, 

that exist as anionic species. 

3.1.5.2. Estimating Metal Distribution in Sediments 

A main objective in terms of assessing the hazard or risk posed by metals in sediments is 

estimating the amount of metal that is potentially bioavailable. The bioavailability of metals in 

sediments is a function of their distribution between the dissolved and solid phases, with 

dissolved metals in porewater generally considered to be the most bioavailable fraction. 

Accordingly, several methods have been developed to estimate the distribution of metals among 

dissolved and solid phases in sediments. These methods have been thoroughly reviewed by 

Mudroch et al. (1999, 1997). Although bioavailability is also a function of aqueous phase 

speciation, limited research has been conducted to estimate metal speciation in pore waters. 

Generally, for the purpose of ecological risk assessments, the exposure of benthic organisms to 

sediment-associated metal is assumed to be proportional to the metal concentrations in interstitial 

water, although some studies indicate that uptake from overlying water (Hare et al., 2003; Roy 

and Hare, 1999) or ingested sediment may be a significant source of body burdens of metals. 

Distribution of metals in sediment pore waters may be determined by field 

measurements, experimental methods, and mathematical modeling, with the latter also requiring 

some field measurements. Concentrations of metals in pore waters may be determined in the 

field by using pore water dialysis chambers or peepers and by methods that separate the solid 

phase from the pore water, although the latter have been shown to be less reliable (Mudroch et 

al., 1997). Several extraction schemes have been developed to determine the distribution of 

metal among operationally defined fractions (e.g., Forstner, 1995; Tessier et al., 1979). 

However, sequential extraction methods do not cleanly distinguish the occurrence and speciation 

of different forms of metals in sediments and soils (Tye et al., 2003; Tack and Verloo, 1999). 

Other experimental methods include leaching tests (e.g., Reuther, 1999). The results of any of 

these methods are concentrations of metals in pore water, which can be related to toxicity 

benchmarks. 

Because of the need to develop Sediment Quality Guidelines (SQGs) for metals that 

explicitly address toxicity and are based on readily measured parameters, several methods have 

been developed. For oxic sediments, Tessier et al. (1993, 1989, 1984; Tessier, 1992) compiled 

partition coefficients of metals that were derived from field studies of freshwater sediments. The 

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partition coefficients are dependent on pH (because Eh is held constant) and are generally linear 

over a range of pore water pH values (see above discussion under ground-water chemistry, 

Section 3.1.4 for the theoretical basis for development of partition coefficients). 

Speciation/complexation models also may be used to estimate fractions of dissolved and bound 

metal species. These models rely on measurements of pH, dissolved oxygen, or Eh to establish 

redox conditions. The models assume that solid-phase binding is governed by sorption to iron 

and manganese oxides. Model estimates are less reliable when other solid-phase substrates are 

dominant (e.g., clay minerals) and are a function of the availability and accuracy of the stability 

constants for the metal-ligand reactions that are used in the calculations. Model estimates are 

less reliable when other solid-phase substrates are dominant (e.g., clay minerals), and they are a 

function of the availability and accuracy of the stability constants for the metal-ligand reactions 

that are used in the calculations. 

For anoxic sediments, the availability of sulfide controls metal distribution and solubility. 

Operationally, AVSs (mainly iron monosulfide) have been considered as a measure of reactive 

sulfides (Forstner, 1995). Studies have demonstrated an inverse relationship between sediment 

acute toxicity and AVSs for marine and freshwater sediments (Di Toro et al., 1992, 1990; 

Ankley et al., 1991) as well as between pore water concentrations and AVSs (Brumbaugh et al., 

1994; Casas and Crecelius, 1994). As a screening-level tool, the acute toxicity of anoxic 

sediments can be assessed by determining the ratio of AVSs to simultaneously extracted metal 

(SEM). Low sediment toxicity is indicated when AVSs are in excess (AVS > SEM), which 

implies sufficient capacity of the AVS to bind essentially all free metal. This topic is further 

discussed in Section 5. 

3.1.6. Soil Chemistry 

The cationic metals occur naturally in soils as oxides and hydroxides (Fe, Mn, Al); to a 

lesser extent they can occur as carbonates, phosphates, and sulfates; and in reducing (usually wet 

or waterlogged) soils as sulfides, which are highly insoluble. The soil parameters important in 

affecting sorption and precipitation reactions and the extent of their influence—and thus 

contaminant bioavailability—depend on the intrinsic properties of the contaminants. In the soil 

environment, metals can exist as cations (having a positive charge), anions (having a negative 

charge), or neutral species (having a zero charge). Their form significantly affects their sorption, 

solubility, and mobility. For example, most soils are chiefly negatively charged; thus, metal 

cations have a higher propensity to be sorbed by soil particles than do metal anions (U.S. EPA, 

2003e). 

Cationic metals can react with inorganic soil constituents (e.g., carbonates, sulfates, 

hydroxides, sulfides), forming aqueous complexes, which may be adsorbed or precipitated in 

mineral form. Most complexation and precipitation reactions are pH dependent (U.S. EPA, 

2003e). 

3-17
As, Cr, Se, and V complex with O and typically exist as anionic species under most 

environmentally relevant scenarios (Bohn et al., 1985; Lindsay, 1979). The most common forms 

of As are arsenate (As(V)) and arsenite (arsenic(III)), which are present in soil solution in the 

form of AsO4

3- and AsO33-, respectively. The chemistry of arsenic resembles that of phosphate 

(Barber, 1995; Bohn et al., 1985). Cr can exist as chromate (Cr(VI) or CrO4

2-), which is usually 

considered more soluble, mobile, and bioavailable than the sparingly soluble chromite (Cr(III)), 

which is normally present in soil as the precipitate Cr(OH)3 (Barnhart, 1997; James et al., 1997). 

Similarly, Se can be present as selenates (SeO4

2-) and selenites (SeO32-). For V, vanadate (VO43-) 

is the most common form (U.S. EPA, 2003e). 

Metals can exist in the pore water as aquo-ions or soluble complexes. The bonding of 

metal species to soil particles can range from ionic to covalent. For most soils in the United 

States, negatively charged sites are more plentiful; less than 5% of the total available charge on 

the soil surface is positively charged. Metals existing as cationic species have a greater 

propensity to associate with such soils. This makes them less bioavailable, but it also results in 

greater loading of metals into the soil ecosystem. Anionic metals generally move into pore 

water—and so are more bioavailable—but leach out of the system much more rapidly. In 

summary, soil pH and availability of charged sites on soil surfaces are the primary soil factors 

controlling release of metals to pore water and, subsequently, bioavailability (U.S. EPA, 2003e). 

3.1.6.1. Key Parameters Affecting Metal Bioavailability in Soils 

From the preceding overview of how the metals and metal compounds interact with soil 

constituents, it is clear that soil plays a very significant role in reducing the potential 

bioavailability of metals in the environment. Given the types of contaminant-soil interactions 

presented, the primary soil factors controlling the potential bioavailability of metals are soil pH, 

the availability and character of sorption sites on soil surfaces, the content of Iron and aluminum 

oxyhydroxides and soil organic matter, and least important, the soil clay mineral content. The 

following discussion briefly details the key soil parameters affecting the various contaminants 

availability to the pore water. 

Soil pH is often termed the master soil variable because it controls virtually all aspects of 

contaminant and biological processes in soil. These processes include dissolution and 

precipitation of metal solid phases, complexation and acid-base reactions of metal species, and 

metal sorption as well as microbial activity. Increasing soil pH also results in an increase in the 

number of negatively charged soil sites, with a concomitant decrease in the positively charged 

sites. Therefore, increasing the soil pH increases the sorption and removal from pore water 

(Bohn et al., 1985). 

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3.1.6.2. Adsorption Behavior of the Metals of Concern 

3.1.6.2.1. Surface area and surface density. In porous media, the most important sorbent solids 

for metals are oxyhydroxides of iron and manganese. Their important surface properties are 

discussed in Langmuir et al. (2003). For a given weight of sorbent, metal sorption capacity is 

proportional to surface area and surface site density. The greatest surface site densities 

(positively or negatively charged sites) are those of organic material and the oxyhydroxides. 

These phases are the strongest and most important sorbents of trace metals. Except for kaolinite, 

the clays (0.02-2 mmol sites/g) have a surface charge that is largely independent of pH, whereas 

the surface charge of organic matter and the oxyhydroxides is strongly pH dependent. 

3.1.6.2.2. Organic matter (organic carbon) content. Organic matter includes plant and animal 

remains in various stages of decomposition, cells and tissues of soil organisms, and substances 

exuded from plant roots and soil microbes (Sumner, 2000). Organic matter is primarily 

composed of carbon, oxygen, and minor amounts of nitrogen (N) and phosphorus (P). On 

average, approximately 58% of organic matter is organic carbon. Organic matter in soils ranges 

from <1% for a sandy soil to almost 100% for a peat soil, with most soils having organic matter 

contents <10% (Bohn et al., 1985). Also, organic matter content is usually higher in surface soils 

or in the root zone and decreases with depth in the soil profile. The organic acid functional 

groups typically present in organic matter have a high affinity for metal cations. 

3.1.6.3. Aging of Metals in Soil 

A distinction should be made between persistence of total metals in soil and persistence 

of bioavailable forms of the metal. As metals age in soils, they decrease in bioavailability. The 

aging process is partially reversible if environmental parameters change (e.g., pH decreases), 

although a portion of the metal ions will be securely entrapped in the soil particle lattice and not 

available to be resolubilized. It has been well documented that metal chemistry in solutions 

freshly added or spiked into soils vary from metal forms in field-contaminated soils. Typically, 

the metal contaminant pool requires time to diffuse into micro- or nanopores and to be absorbed 

into organic matter and soil particles. These slow reactions are attributed to micropore diffusion, 

occlusion in solid phases by (co)precipitation and (co)flocculation, and cavity entrapment. 

Although the slow reactions play a key role in metal bioavailability, their rates, mechanisms, and 

controlling factors have not been comprehensively elucidated. Evidence of aging processes is 

provided by studies of metal extractability and lability (Tye et al., 2003; Hamon et al., 1998). It 

has been frequently observed that easily extractable pools revert with time (~1 year) to more 

strongly bound forms. Isotopic dilution provides a useful way to quantify changes associated 

with progressive attenuation of metals in soil. After 1 year aging reactions are almost complete 

and are mostly reversible. At present, information regarding the aging reactions of different 

3-19
metals and metalloids and sorbing solids, is very limited, so it is not possible to generalize which 

metal(s) age the fastest or with greater/less reversibility. 

3.1.6.4. Dissolution and Transformation of Metals 

The dissolution and transformation of a metal compound in soil is related to a series of 

chemical and physical properties characteristic of the compound itself and of the soil. 

Environmental parameters such as temperature and humidity have a strong influence on the rate 

of transformation. When assessing the transformation of a compound in soil, the risk assessor 

should keep in mind that aging reactions may take place at the same time as transformation and 

dissolution. When metal salts are added to soil, the form of the salt dictates the rate and amount 

of soluble metal that will form in the pore water. Insoluble forms of metals (e.g., vanadium 

pentoxide [V2O5]) will transform to soluble free ion (V) at a slower rate than will soluble metal 

salts (e.g., Na3VO4). However, the rate of formation of the free ion is not proportional to the 

dissolution rate of the salt because aging reactions will remove the free ion from the pore water. 

The risk assessor should consider the relative rates of dissolution and aging simultaneously to 

accurately predict pore water concentrations. 

3.1.6.5. Soil Metal Transfer to Plants 

The “soil-plant barrier” concept was introduced to communicate how metal addition, soil 

chemistry, and plant chemistry affect risk to animals from metals mixed in soil (Chaney, 1980). 

Reactions and processes that take place at the soil-plant barrier are influenced by the following 

factors: (1) soil solid phases (e.g., iron, aluminum, and manganese oxyhydroxides and organic 

matter) may have adsorptive surfaces that influence soil chemistry; (2) adsorption or 

precipitation of metals in soils or in roots limits uptake-translocation of most elements to shoots; 

and (3) the phytotoxicity of Zn, Cu, Ni, Mn, As, Bo, Al, and F, for example, limits residues of 

these elements in plant shoots. More recently, inductively coupled plasma-mass spectrometry 

(ICP-MS) and other very sensitive analytical methods have been used to examine soil solution 

and soil-plant transfer of 60 elements as a function of soil pH (Tyler and Olsson, 2001a, b). 

These studies provide evidence that further supports the concept of a soil-plant barrier 

Several areas for potential errors in the research methodology should be avoided when 

making these comparisons: 

• First, comparison of application rates is valid only after the system has been 

equilibrated using accepted methods. 

• Second, soil pH levels should be equal across rates studied; co-variance of soil pH 

should be used to correct for unequal soil pH (Bell et al., 1988). 

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• Third, the metal concentration in the source applied affects the slope of metal uptake: 

higher metal concentration in the source means higher phytoavailability at equal 

metal applications (Jing and Logan, 1992). 

Strongly acidic soils increase plant uptake of Zn, Cd, Ni, Mn, and Co and increase the 

potential for phytotoxicity from Cu, Zn, and Ni. In alkaline soils, the high pH increases uptake 

of Mo and Se. Pb and Cr are not absorbed by plants to any significant extent at any pH (Chaney 

and Ryan, 1993). However, each element should be considered separately because of its unique 

chemistry. For example, arsenate is more strongly adsorbed than is arsenite; when a soil is 

flooded to grow rice, soil microbes can reduce arsenate to arsenite, and the higher concentration 

of dissolved arsenite can be phytotoxic to rice in more highly contaminated soils. Most other 

elements have little potential for redox change with change in the redox status of soils. Reduced 

soils can form sulfide, and sulfide forms low-solubility compounds with most of the metals of 

concern in soils, including Pb, Zn, Cd, Cu, and Ni (see above discussion on sediment chemistry 

Section 3.1.5). For essential elements (e.g., Zn, Cu, Ni), low-solubility species can result in 

deficiency syndromes. Upon oxidation of the soil, sulfide is quickly oxidized, and the metals are 

returned to more normal equilibrium reactions of aerobic soils. 

3.1.7. Atmospheric Behavior/Chemistry 

Most metals and metal compounds exist in the solid phase under normal environmental 

conditions and thus occur almost exclusively in the particle phase of the atmosphere, where they 

are ubiquitous. Anthropogenic sources include combustion from fossil fuels and metal industries 

as well as industrial sources employing specific metal compounds in specific processes. Some 

airborne metals (e.g., manganese and nickel) may derive largely from crustal sources (U.S. EPA, 

1996a). Richardson (2002) included volcanic eruptions and emissions, entrainment of soil and 

dust, entrainment of sea salt spray, and natural forest fires as significant natural metals emission 

sources. 

For purposes of health risk assessment, particle size is important. The aerodynamic size 

and associated composition of particles determine their behavior in the human respiratory 

system. Furthermore, particle size is one of the most important parameters in determining the 

atmospheric lifetime of particles, which may be a key consideration in assessing inhalation 

exposures as well as exposures related to exposure pathways involving deposition onto soil or 

water (U.S. EPA, 1996a). Metals emitted by combustion processes (e.g., the burning of fossil 

fuels or wastes) generally occur in small particles or fine fraction, which is often characterized 

by particles less than 2.5 microns in diameter (PM2.5). In contrast, the larger sized, coarse mode 

particles result from mechanical disruption, such as crushing, grinding, and evaporation of 

sprays, or suspensions of dust from construction and agricultural operations. Accordingly, 

metals in coarse mode particles (i.e., those larger than approximately 1–3 μm) are primarily those 

3-21
of crustal origin, such as Silicon (Si), Al, and Fe (U.S. EPA, 1996a). The fine versus coarse 

distinction simply differentiates two relatively distinct size distributions of particles, the 

separation point of which occurs in the range of 1 to 3 μm. The distinction does not refer 

directly to particle sampling methods or size fractionations particular to health risk assessment 

(U.S. EPA, 1996a). 

Fine and coarse particles typically exhibit different behavior in the atmosphere; fine 

mode particles exhibit longer atmosphere lifetimes (i.e., days to weeks) than coarse particles and 

tend to be more uniformly dispersed across a large geographic region (U.S. EPA, 1996a). 

Relatively lower dry deposition velocities of fine particles contribute to their persistence and 

uniformity throughout an air mass (U.S. EPA, 1997c). The larger coarse particles (i.e., greater 

than 10 μm) tend to rapidly fall out of the air and have atmospheric lifetimes on the order of 

minutes to hours, depending on their size and other factors (U.S. EPA, 1996a). 

In most cases, metals do not undergo transformation while in the particle phase; thus, 

their removal from the atmosphere is regulated by the rate at which the particles themselves 

participate in wet and dry deposition processes. For example, metals such as As, Be, Cd, and Pb 

are generally found in airborne compounds with a single predominate oxidation state (As(III), 

Be(II), Cd(II), Pb(II)). Some metals (e.g., the transition metals Cr, Mn, and Ni) present the 

possibility of changing oxidation state in situ in the particle, although little is known of these 

processes (U.S. EPA, 2003d). This is an important consideration for health risk assessment as 

the different oxidation states also differ in toxicity (such as for Cr). 

For metals that can change oxidation states, much of the atmospheric chemistry takes 

place in the aqueous phase, such as cloud droplets or water films on particles. Metal salts and 

oxides that dissolve in water can undergo several reversible reactions, including hydration, 

hydrolysis, polymerization, and reaction with other anions. The equilibrium between these 

forms depends on the atmospheric conditions, the equilibrium and solubility constants, and the 

concentrations of other chemicals. Transformations between oxidation states can occur either to 

increase the oxidation state (such as oxidizing Cr(III) to Cr(VI)) or to reduce it. These oxidation 

or reduction reactions can occur through reaction with other species, such as dissolved metals, 

reduced sulfur species, and organic compounds (Seigneur and Constantinou, 1995). Although 

models exist that can be used to estimate metal speciation in aerosols with liquid water, the 

reactions are still highly uncertain. 

Hg is an exception among the commonly occurring metals; it exists primarily in the vapor 

phase under most conditions but can also occur in particle and aqueous phases. The risk assessor 

should consider at least three species of Hg: elemental Hg, which is largely present as a gas; 

divalent (Hg(II)) inorganic Hg compounds, which are more water soluble; and particulate-phase 

Hg (Schroeder and Munthe, 1998; U.S. EPA, 1997b). The behavior of Hg in the atmosphere 

depends strongly on the oxidation state. Elemental Hg is capable of being transported long 

distances, even globally; divalent Hg deposits within a few hundred kilometers of sources; and 

3-22
particulate mercury is deposited at intermediate distances, depending on the particle size 

(Schroeder and Munthe, 1998). Elemental Hg that is deposited can be reemitted from the 

surface, as can divalent and particulate Hg after biological or chemical reduction to the elemental 

form. 

In the gas phase, elemental Hg can be oxidized to divalent Hg by O3, OH, H2O2, and 

molecular chlorine (Cl), although other halogen atoms also might be important (Schroeder and 

Munthe, 1998). In the aqueous phase, elemental Hg can be oxidized by OH, O3, and dissolved 

chlorine, and divalent mercury can be reduced by processes such as reaction with HO2 and 

S(IV). Both vapor-phase and aqueous atmospheric chemistry may involve multiple phases. 

EPA has made a substantial effort to evaluate the atmospheric fate of Hg as a result of the 

requirements of the Clean Air Act. EPA (1997b) contains a comprehensive evaluation of 

mercury’s atmospheric fate, but this is an area of ongoing research and controversy. EPA 

continues to actively investigate Hg behavior in the atmosphere (e.g., Landis and Stevens, 2003; 

Jaffe et al., 2003; Bullock and Brehme, 2002; U.S. EPA, 2001a; Bullock, 2000a, b). 

Most sampling and analytical techniques published by EPA for metals in air are oriented 

toward evaluation of particular-phase total metals rather than metal species (U.S. EPA, 1999a). 

These methods involve collection of a particular size fraction of particles (e.g., PM2.5, PM10, and 

total suspended particulates), with subsequent analysis by x-ray fluorescence, atomic absorption, 

inductively coupled plasma, proton-induced x-ray emission, or neutron activation analysis 

gamma spectroscopy techniques. The one notable exception is a method for Hg(Method IO-5) 

that speciates vapor and particulate forms. To the extent that metals are sorbed to particulate 

phases, analysis of individual metal species can, at least theoretically, be accomplished by the 

same techniques used to analyze those species in other solid media. 

Potential for Inhalation 

Inhalation of metals is addressed in Section 4.2.5.1. Particle size is likely to be the most 

important determinant of potential exposure. Particle size also influences the overall fate of 

metals in the atmosphere. 

3.2. METAL FATE AND TRANSPORT 

Major metal sources to water and land include diverse manufacturing, mining, 

combustion, and pesticide activities. Major atmospheric sources are oil and coal combustion, 

mining and smelting, steel and iron manufacturing, waste incineration, phosphate fertilizers, 

cement production, and wood combustion (Haygarth and Jones, 1992). Metals from these 

atmospheric sources can find their way into soils, sediments, and water. Other major sources to 

aquatic and terrestrial systems include chloralkali, acid, pigment, electronics, and copper sulfate 

production. 

3-23
Some exposure assessments do not involve anthropogenic releases of metals to the 

environment. Rather, they focus on changes in exposure to environmental metals that result 

from other aspects of human activities. For example, acidification of freshwater changes Al 

phase association and speciation, resulting in an increased dose of naturally occurring metals to 

aquatic biota (e.g., Campbell et al., 1992). Intensive irrigation mobilizes Se that is naturally 

present in relatively high concentrations in western soils, and consequent evaporative 

concentration in wetlands, impoundments, and other low-lying areas in arid regions of the U.S. 

can lead to toxic exposures (e.g., Wu et al., 1995). 

3.2.1. Aquatic and Terrestrial Transport Pathways for Metals 

When an exposure assessment is completed for a site, the available data are usually 

limited in their coverage with respect to the characterization of exposure levels over both time 

and space. Depending on the situation, it may be advantageous to use a transport and fate model 

(i.e., a computational model) to fill in the data gaps, such that an improved characterization of 

exposure levels is available for use by the risk assessor. Models are also useful in situations 

where it is desirable to have an estimate of future exposure levels that are expected to result from 

the implementation of remediation measures. These results can be used to quantitatively 

evaluate the effectiveness of alternative remediation scenarios that are being considered. The 

models can also be used to refine the design of the viable alternatives so that the risk assessor can 

develop an optimal remediation strategy. 

Although numerous models are available for use, most are based on the same two 

fundamental principles: (1) metals are ubiquitous in the environment, as they are found in the 

aquatic, terrestrial, and atmospheric compartments; (2) within each compartment, they are 

present in association with water (freely dissolved metal or as organic and inorganic metal 

complexes), particles (sorbed, precipitated, or incorporated within a mineral phase), and air. The 

evaluation of metal transport therefore requires evaluation of the distribution of the metal among 

these phases, within each compartment (air, water, soil), as well as the movement (i.e., the 

transport) of each of these within and among the various compartments. It is important for the 

assessor’s model to simulate the movement of water and particles explicitly because this 

provides a way to evaluate differences in the degree to which various chemicals/metals may be 

transported in association with particles (i.e., via settling and resuspension) or in association with 

the dissolved phase (diffusive flux of dissolved metal). The analyst represents the environmental 

setting of interest as a series of discrete, interconnected volumes. Mass balance equations for air, 

water, solids, and metal are then formulated for each volume to obtain a system of mass balance 

equations that may then be solved for the concentrations of interest over both time and space. 

Including all compartments in every model is not necessary. For example, models for a site that 

is impacted by a smelter might call for use of a model of an atmospheric compartment (e.g., to 

simulate transport of a release from a stack) and a terrestrial compartment (to simulate fate of 

3-24
atmospheric inputs to the soil). Alternatively, for an aquatic setting dominated by previously 

contaminated sediments, it may be sufficient to consider water and sediment transport alone and 

to neglect the atmospheric and terrestrial compartments. 

Although the underlying principles of most models are similar, the included features vary 

widely from one model to the next. The output from a relatively simple model may be adequate 

for decision-making purposes in some instances. Some relatively simple models are limited in 

their applicability to steady-state analyses and spatially uniform conditions, but if this will 

provide a reasonable and/or conservative representation of conditions at a site, it may provide the 

analyst with a useful and cost-effective modeling alternative. In other cases, the analysis may 

require the completion of time-variable simulations to properly represent conditions that vary 

over time, such as daily or seasonal variations in flow and upstream boundary concentrations, 

point source loads, and pulse exposures. The analyst should select an appropriate model because 

not all models will be applicable to every situation. Although, in principle, the more 

sophisticated models provide the risk assessor with the capability to complete a more detailed 

and mechanistically based analysis than will a simple model, successful application of these 

models will require greater resources (data, time, and funding) than will the use of a simpler 

model. The analyst must also possess a relatively high level of modeling expertise to use the 

more sophisticated modeling approaches. It is for this reason that the more sophisticated models 

are usually reserved for use in higher level, definitive assessments. 

Many of the models available for use in evaluating the transport and fate of metals were 

originally developed for application to neutral organic chemicals. As a result, these models 

frequently include a variety of reactions that are not necessarily relevant to an analysis of metal 

transport and fate (e.g., biodegradation, photo-oxidation, and volatilization). Although these 

models still may be of use in an exposure assessment for metals (the nonapplicable processes 

often may be bypassed), a more significant problem is that they often fail to represent some 

important metal-specific processes. For example, the evaluation of metal speciation and metal 

partitioning between dissolved and particulate phases will be represented only in a very simple 

manner in such models. This limitation may be overcome, at least in part, by performing the 

requisite metal-specific analyses with a stand-alone chemical equilibrium model, but this 

approach will place an added burden on the analyst to integrate the results of the two models in a 

technically defensible manner. 

Although models that include some metal-specific capabilities will be discussed as part of 

this Framework, no single model that is currently available for use includes all the metal-specific 

features that would be desirable for use. More detailed discussions of these and other transport 

and fate models, as well as a number of chemical equilibrium models, may be found in Paquin et 

al. (2003) for aquatic systems and Allen (2002) for terrestrial systems. Additionally, work is in 

progress to develop updated models that will offer improved metal-specific capabilities (e.g., the 

3-25
Unit World model (Di Toro et al., 2001). As a result, metal fate and transport models should be 

viewed as an evolving tool, and new models may be expected in the near future. 

3.2.1.1. Aquatic Transport Models 

Modeling of metal transport and fate within aquatic systems involves the representation 

of hydrodynamic transport to simulate movement of water, particulate transport to simulate the 

movement of particles, and chemical transfers and kinetics to simulate exchange of metal 

between dissolved and particulate phases and between the water column and benthic sediment 

(Figure 3-2). The risk assessor has the option of using independent hydrodynamic transport, 

sediment transport and chemical fate models, or an integrated model that incorporates all these 

processes. 

Figure 3-2. A generalized model framework for chemical fate and transport 

in an aquatic system. 

Source: Paquin et al. (2003). 

Modeling the movement of metals through an aquatic system begins with a 

characterization of the water movement through the system. The time scale for the 

hydrodynamic analysis should be represented in a way that will satisfy the needs of the sediment 

transport and chemical fate analyses that are also being performed. For example, low-flow 

3-26
conditions associated with minimum dilution may be judged to 

be the most critical conditions in a setting involving a point 

source discharge, while peak flow conditions may need to be 

simulated in a setting where resuspension of contaminated 

sediments is the primary concern. A steady-state model might 

be appropriate for use in the former case, and a time-variable 

model would likely be needed in the latter case. Thus, the 

details of the specific problem setting will necessarily have an influence on both model selection 

and how the model will be used. 

The risk assessor also should be aware of several reviews of fate and transport models 

that have been completed since the early 1980s. In earlier reviews, Delos et al. (1984) reviewed 

models for wastewater load allocations, and Mills et al. (1985) described screening-level 

analyses for water quality assessments of conventional pollutants. The Agency subsequently 

prepared an updated summary of the features included in the water quality models that were 

available for use in exposure assessments (U.S. EPA, 1987). Schnoor et al. (1987) 

simultaneously published their review and included detailed descriptions of fate and transport 

models and their required input parameters. Later, EPA published a review describing the use of 

modeling tools for the development of Total Maximum Daily Loads (TMDLs) in watersheds 

(U.S. EPA, 1997a). More recently, Paquin et al. (2003) completed a review of exposure, 

bioaccumulation, and toxicity models for aquatic systems, with a focus on their applicability to 

metals (exclusive of organometallics). Because of the advances that have been made since the 

mid-1990s with regard to the development of fate and transport models, including some recent 

efforts to couple these models with metal speciation models and more sophisticated stand-alone 

hydrodynamic and sediment transport models, the latter reviews by EPA (1997a) and Paquin et 

al. (2003) tend to include the most up-to-date information with regard to the availability of 

models that are appropriate for use. These reviews also include example applications of many of 

the models discussed. 

Fate and transport analyses may be performed by using an integrated hydrodynamic, 

sediment, and chemical transport model or by employing what tend to be relatively sophisticated 

stand-alone versions of these three submodels. The advantage of the former approach is that 

integration of the hydrodynamic, sediment, and chemical transport results takes place in a 

seamless manner with limited need for an assessor’s intervention. This is in contrast to the use 

of stand-alone models, where the output of one model should be formatted in a way that ensures 

it is amenable to use with the other models that are to be applied. A distinct advantage of the 

latter approach is that it has the potential to reduce the time needed to complete a model run, an 

important consideration for a multiyear simulation of a large and complex problem setting. For 

example, it may not be necessary to repeat the simulation of hydrodynamic and sediment 

transport if the model input being modified affects only chemical transport (e.g., partition 

Integrated Models 

The fate and transport of 

metals in aquatic systems is 

most reliably predicted using 

integrated models rather than 

stand-alone hydrodynamic or 

sediment transport models. 

3-27
coefficient). A similar line of reasoning would apply to the use of an integrated transport model 

that includes metal speciation versus one that uses a stand-alone chemical equilibrium model to 

satisfy this need. 

The partition coefficient, which controls the 

distribution of metal between the dissolved and particulate Analytical Solution Models 

phases, is considered to be a key model parameter in Analytical solution models 

of aquatic system fate and 

chemical/metal fate and transport evaluations (see Section transport are the simplest 

3.1.4). It is important for several reasons. First, the models with the lowest 

computational requirements. 

distribution of metal between the dissolved and particulate They are solved analytically to 

provide concentrations over phases has a direct impact on the magnitude of particulate metal of time or space. 

fluxes of metal that occur in association with the settling and 

resuspension of sorbed metal. Second, the partition coefficient also controls the magnitude of 

diffusive fluxes of metal between the sediment interstitial water and the overlying water column, 

as this flux is proportional to the concentration gradient of total dissolved metal (free ionic metal 

+ metal-DOC and metal-inorganic ligand complexes) between these compartments. Partitioning 

reactions also affect the metal bioavailability and the route of exposure to a metal (food vs. 

water). Given its importance, the risk assessor should recognize that the assumption of 

equilibrium partitioning is a potential oversimplification in some situations (e.g., near a point 

source discharge to a receiving water or immediately following a pulse exposure, such as an 

overflow from a storm sewer system). Under these conditions, the assumption of equilibrium 

between the dissolved and particulate phases may be less appropriate than is otherwise the case 

and additional uncertainty will necessarily be associated with the model results. If this situation 

exists, consideration should be given to conducting special studies (e.g., measurement of free 

metal concentrations over time in a water sample) to test the validity of the equilibrium 

assumption. 

Modeling of Hg and metalloids, such as As and Se, is complicated by transformation 

processes that change the form of the metal. Similarly, metals that readily change oxidation state 

(e.g., Cr) also require additional considerations in fate modeling. Many of the same transport 

models can be used, but input parameters will require modification. Such models are beyond the 

scope of this framework, which is focused on inorganic forms of metals. See, for example, the 

Hg cycling model by Hudson et al. (1994) for further guidance on fate and transport models for 

Hg. All modeling exercises are limited by the validity of the model Framework, the accuracy of 

input parameters, and the experience of the analyst (Dzombak and Ali, 1993). Therefore, all 

exposure assessments should include an explicit description of model assumptions and 

associated uncertainties. 

3-28
3.2.2. Terrestrial Fate and Transport 

Movement of metals through soils is dependent on the chemical properties controlling 

speciation, the presence of ligands that control complexation of metals within pore water (and 

ground water) and adsorption onto mineral surfaces, and the rate of water flux through the soil. 

Metals are lost from the soil by leaching into ground water and through metal specific uptake by 

plants (see Section 3.1.6.5). Sections 3.1.4 and 3.1.6 review the processes and models that 

predict movement of chemicals through soils or partitioning onto mineral surfaces (i.e., partition 

coefficients). These serve the same role as the fate and transport models discussed above for 

aquatic systems and therefore are not repeated here. 

3.3. ATMOSPHERIC FATE AND TRANSPORT 

Fine and coarse particles typically exhibit different behavior in the atmosphere: fine 

mode particles exhibit longer atmospheric lifetimes (i.e., days to weeks) than coarse particles and 

tend to be more uniformly dispersed across a large geographic region (U.S. EPA, 1996a). 

Relatively lower dry deposition velocities of fine particles contribute to their persistence and 

uniformity throughout an air mass (U.S. EPA, 1997c). The larger coarse particles (i.e., greater 

than 10 μm) tend to rapidly fall out of the air and have atmospheric lifetimes on the order of 

minutes to hours, depending on their size and other factors (U.S. EPA, 1996a). 

In the atmosphere, the behavior of Hg depends strongly on its oxidation state. Elemental 

Hg is capable of being transported long distances, even globally; divalent Hg deposits within a 

few hundred kilometers of sources; and particulate Hg is deposited at intermediate distances, 

depending on the particle size (Schroeder and Munthe, 1998). Elemental Hg that is deposited 

can be reemitted from the surface, as can divalent and particulate Hg after biological or chemical 

reduction to the elemental form. 

Deposition 

Because most metals occur as particulate matter (hereinafter referred to as particulates), 

their removal from the atmosphere is governed by the rates of wet and dry deposition. These 

deposition processes represent an important route of exposure for plants, animals, and humans. 

Some of the deposited metals can come to reside on the surfaces of vegetation, resulting in 

potential exposure to the plants and to animals that eat the plants. Metals deposited from the 

atmosphere also come to reside in the leaf litter and uppermost soil surfaces, where they can be a 

source of exposure for soil organisms, people, and wildlife. 

3-29
4. HUMAN HEALTH RISK ASSESSMENT FOR METALS 

Source: IPCS-EHC #214(2000). 

Human Health Risk Assessment Paradigm 

Source: IPCS-EHC #214(2000).

The National Research Human Health Risk Assessment Paradigm

Council (NAS/NRC, 1996, 1994b, 

1983), of the National Academy of 

Sciences (NAS), described four 

phases to the human health risk 

assessment paradigm (Hazard 

Identification, Dose-response 

Assessment, Exposure Assessment, 

and Risk Characterization) and 

identified risk communication as a 

fifth area of study. These principles 

have been further addressed in 

EPA’s Risk Characterization 

Handbook (U.S. EPA, 2000c). In brief, hazard identification (referred to as “hazard 

characterization” in recent EPA documents) involves the determination of whether a chemical is 

or is not causally linked to particular health effects. Dose-response involves the determination of 

the relationship between the magnitude of exposure and the probability of occurrence of the 

health effects in question. A parallel step in the process toward the hazard identification and 

dose-response assessment is exposure assessment. In Exposure Assessment, the risk assessor 

quantifies the total exposure to a toxic agent in the environment based on amount taken into the 

body, including any combination of the oral, inhalation, and dermal routes of exposure. For 

some assessments specific to a single exposure route, exposure may be expressed as an 

environmental concentration (e.g., ambient air or water concentrations). Depending on the 

application, the exposure assessment may be specific to a site, a population at a specific location, 

or it may broadly cover a region or an entire nation. Risk Characterization is the final step in the 

NAS paradigm. In this phase, the risk assessor summarizes and interprets the information from 

hazard identification, dose-response, and exposure steps, often by quantitatively comparing 

exposures with doses that are associated with potential health effects. Risk Characterization 

addresses the nature and often the magnitude of the human health risks, including attendant 

uncertainty. These steps are addressed in greater detail in the following sections, with particular 

attention to the aspects specific to metals. 

The information provided here complements that given by the available Agency guidance 

for the risk assessment process, e.g., for carcinogen risk assessment (U.S. EPA, 2005, 1986), 

exposure assessment (U.S. EPA, 1992c), developmental toxicity (U.S. EPA, 1991), neurotoxicity 

(U.S. EPA, 1998c), chemical mixtures (U.S. EPA, 2000b, 1986), and cumulative risk (U.S. EPA, 

4-1
2003e), and focuses on the unique and specific characteristics of metals and metal compounds 

that may be applied in metals risk assessments for human health. 

4.1. METALS PRINCIPLES 

Metals are associated with a variety of health effects that are reviewed in detail in EPA’s 

Integrated Risk Information System (IRIS) Toxicology Reviews, the Agency for Toxic 

Substances and Disease Registry (ATSDR) Toxicological Profiles, the World Health 

Organization’s International Programme for Chemical Safety (WHO/IPCS) Environmental 

Health Criteria Documents, and metal toxicology reviews (e.g., Lukaski, 1999; Hathcock, 1996; 

Mertz, 1995, 1993; Wallach, 1985). Metals have specific attributes that should be considered in 

all risk assessments. These principles for metals risk assessment (see Chapters 1 and 2) apply in 

various ways to human health risk assessments, depending on the scale of the assessment (site 

specific, regional, or national). This section describes applications of the metals principles to 

human health assessments within the standard risk assessment framework. Specifically, they fall 

into the risk assessment paradigm as follows: 

Background levels Exposure Assessment 

Mixtures Exposure and Effects Assessment 

Essentiality Effects Assessment 

Forms of metals Exposure Assessment (bioavailability) and 

Toxicokinetics/toxicodynamics Effects Assessment (absorption, distribution, 

metabolism, and excretion [ADME] and 

toxicity) 

Often times, human health risk assessors start their analysis with a metal-specific reference 

value (RfD/RfC) and/or cancer potency factor that has been developed through a process separate 

from the health risk assessment. The role of the human risk assessor is to appropriately integrate the 

reference values and potency factors with the exposure assessment. Thus, the risk assessor needs an 

understanding of the toxicological endpoints and mechanisms of action that underlie the derivation of 

these values to ensure that, for example, the appropriate population and life stages are addressed, 

appropriate dietary aspects are taken into consideration, and the appropriate exposure pathways are 

considered. For metals, frequency and duration of exposure, as well as exposure concentrations and 

metal species, are important parameters for the risk assessor to consider for accurate dose 

assessments. 

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4.2. HUMAN EXPOSURE ASSESSMENT 

NAS/NRC (1994b) defines 

exposure assessment as “the 

determination of the intensity, 

frequency, and duration of actual or

hypothetical exposures of humans to 

the agent in question. In general, 

concentrations of the substance can 

be estimated at various points from 

its source.” Although there is no 

specific guidance exclusively for 

metal exposure assessment, EPA h

published guidelines for exposure 

assessment (U.S. EPA, 1992c), exposure factors (U.S. EPA, 1997e, 1989a), exposure factors for 

children (U.S. EPA, 2002b), and assessment of early lifestage exposure (U.S. EPA, 2005). 

Additional reports and principles have been published by the Centers for Disease Control and 

Prevention (CDC, 2003) and IPCS (2000, 1999, 1994b). 

as 

Critical Factors Influencing Interaction Between Host and Chemical 

Source: IPCS-EHC #155 (1994). 

Assessment of human exposures to metals, as with any chemical agent, includes: 

(1) identifying how people come into contact with metals in the environment 

(2) determining the concentrations of specific forms (speciation) of the metal in specific 

medial (e.g., soil, water, air, and biota) 

(3) identifying the pertinent exposure metric (via consideration of dose-response 

assessment) 

(4) estimating the exposure metric (e.g., oral intake, inhalation exposure concentration, 

blood concentration), which may involve quantifying relationships between exposure 

concentrations and intakes and 

(5) identifying sources of uncertainty and natural variability and, where possible, 

quantifying these in estimates of exposure. 

4.2.1. Background Levels 

Metals are naturally occurring constituents in 

the environment. As a result of industrialization, 

current environmental levels of metals can be elevated 

relative to naturally occurring levels. This may result 

in a wide variability in the intake of some metals in 

food (e.g., seafood), drinking water, or air. Strategies 

for estimating metal concentrations in air, soil, and 

4-3

Background Metal Concentration 

As a result of industrialization, current 

environmental levels of metals can be 

elevated relative to naturally occurring 

levels. Depending on the purpose of 

the risk assessment, assessors should 

distinguish among naturally occurring 

levels, existing background levels, and 

contributions from specific activities at 

the local or regional level. 
water are discussed in Section 3. In human health risk assessments, the term “background” 

refers to all existing metal sources except the targeted source. A particular challenge for the risk 

assessor may be assessing the metal levels associated with the source(s) of interest in light of 

levels derived from natural and other anthropogenic sources. 

4.2.2. Bioavailability 

The term “environmentally available fraction” refers to the portion of total metal in soil, 

sediment, water, or air that is available for physical, chemical, and biological modifying 

influences (Lanno, 2003) and represents the total pool of metal at a given time in a system that is 

potentially able to contact or enter an organism. Of that portion, the bioaccessible fraction (BF) 

is the amount that actually interacts at the organism’s contact surface and is potentially available 

for absorption or adsorption (if bioactive upon contact) by the organism. Bioavailability is the 

extent to which bioaccessible metals (see Section 2.3) cross biological membranes, expressed as 

a fraction of the total amount of metal the organism is proximately exposed to (at the sorption 

surface) during a given time and under defined conditions. 

The concept of metal bioavailability includes metal species that are bioaccessible and are 

absorbed or adsorbed (if bioactive upon contact) with the potential for distribution, metabolism, 

elimination, and bioaccumulation in the organism. Metal bioavailability is specific to the metal 

salt and particulate size, the receptor and its specific pathophysiological characteristics, the route 

of entry, duration and frequency of exposure, dose, and the exposure matrix. The metal salt is 

influenced by properties of the environment such as pH, particle size, moisture, redox potential, 

organic matter, cation exchange capacity, and acid volatile sulfides. Depending upon the 

assessment, it may be appropriate for the risk assessor to consider metal bioavailability and 

compare the bioavailable fractions used to estimate reference doses (RfDs), or the equivalent, to 

those measured in the diet, drinking water, or air. 

Prediction of toxicity due to exposure to inorganic metals is complicated by wide 

variations in the bioaccessibility and bioavailability of accumulated metals. The form of the 

metal (chemical species, compound, matrix, and particle size) influences its bioaccessibility. In 

turn, the metal form is impacted by properties of the 

environment such as pH, particle size, moisture, redox Bioavailability/Hazard 

potential, organic matter, cation exchange capacity, and acid Relationship 

volatile sulfides. Bioavailability (how much of the ingested If two substances were equally 

toxic at comparable levels of target 

metal interacts at the target site) is influenced by nutritional organ exposure, the substance 

state (deficiency versus excess), age, sex, physiological state, with the higher intrinsic 

bioavailability would pose the pathological conditions, and interactions with other greatest risk. 

substances present. 

It is important that the Exposure Analysis describes the same bioavailable fraction of the 

metal(s) of concern as that used when estimating the reference value (e.g., the RfD). For 

4-4
example, measuring total metals in dietary items may include insoluble forms of the metal 

(particularly if soil contamination is present on the foodstuff), so effective exposure is 

overestimated. There are both direct and indirect approaches to address the relative 

bioavailability of metals in the environment: (1) conduct new animal toxicology studies using the 

metal form encountered in the site assessment; (2) use adjunct scientific data to derive an 

adjustment to the effective dose identified in the animal study (e.g., data on the distribution of 

chemical forms of the metal in the environment or at a contaminated site); or (3) use a default 

assumption that the metal in the environmental samples is the same as that tested. Of the three 

approaches, the first is more scientifically sound. The second option might be available in some 

circumstances but is usually precluded by time and financial resource limitations, and the third 

option, is the most health-conservative. 

A fourth alternative conducted for site-specific assessments is for the risk assessor to 

estimate bioavailability through solubility studies or limited bioavailability studies of specific 

samples from the site. For example, arsenic bioavailability has been estimated for soils from 

various contaminated sites (Ng et al., 1998; Freeman et al., 1995, 1993) and also through a series 

of solubility studies of soil from a site contaminated with mine tailings (Ng et al., 1998; Salocks 

et al., 1996). Additional examples are animal feeding studies with juvenile swine for lead 

bioavailability adjustments or in vitro tests, although the Agency currently requires additional 

validation of the latter approaches before they can be used as the sole basis for making 

bioavailability adjustments (U.S. EPA, 2006a). 

4.2.3. Susceptible Populations 

Risk assessors must specifically consider population subgroups, which may have a 

greater risk to metals than the general population (U.S. EPA, 2006b). Factors influencing 

susceptibility to metals include life stage, life style, gender, reproductive status, nutritional state, 

pre-existing health conditions or disease, and genetic polymorphisms. Children and elderly 

people do not regulate metal uptake and distribution efficiently and may be at higher risk of 

accumulating toxic levels (U.S. EPA, 2006b). Pregnant and lactating women have a higher 

demand for essential elements, and lack of adequate levels of protein in the diet can affect the 

bioavailability of essential elements (NAS/NRC, 2000). Individuals with chronic liver or kidney 

disease may have a lower threshold for effects because these are two of the major target organs 

of metal toxicity. Several well-known, heritable genetic alterations affect people’s ability to 

regulate Cu or Fe, resulting in various deficiency or toxicity problems (WHO/IPCS, 2002). 

Although many of these same factors are considered in all human health risk assessments, each 

has attributes specific to metals-associated risks. 

4-5
4.2.3.1. Life Stage 

In addition to higher intake per kilogram of body weight (Plunkett et al., 1992), children 

may also be more sensitive than adults to metal irritants since sensitivity to skin irritants is 

generally considered to decrease with age. Infants in the immediate postnatal period can also be 

more susceptible to systemic effects of metals because absorption of essential metals is poorly 

regulated (WHO, 1996). At the other extreme, older adults are more sensitive to metals that 

target the kidney (e.g., Cd ) because renal function declines with age. Efficiency of intestinal 

uptake of some trace metals, particularly Zn and Cu, also declines as people age (WHO, 1996; 

IPCS, 1994). 

4.2.3.2. Demographics 

Differences in lifestyle influence metal exposure. The risk assessor should explicitly 

investigate different lifestyles of the population of concern. For example, the use of dietary 

supplements and other consumer products containing essential elements has increased. In 

addition, folk remedies such as colloidal silver “cure-alls” and folk remedies containing lead 

tetroxide may contain high levels of metals (McKinney, 1999; Yanez et al., 1994; Pontifex and 

Garg, 1985; Trotter, 1985; Bose et al., 1983; CDC, 1983, 1982, 1981; Geffner and Sandler, 

1980). Smoking provides significant exposure to some metals (e.g., Cd) and can potentiate the 

effects of exposures from other sources, and excessive alcohol consumption can exacerbate 

metal effects. 

4.2.3.3. Pregnancy and Lactation 

Pregnancy and lactation increase demand for some essential metals, particularly Cu, Zn, 

and Fe (NAS/IOM, 2003; Picciano, 1996). Because of physiological changes that include higher 

Fe and Ca requirements, hormonal changes, and susceptibility to respiratory disease, Zuurbier 

and Van den Hazel (2005) suggested that pregnant women could be predisposed to the toxic 

effects of beryllium (Be), Pb, and Mn (2005). Recommended dietary allowances (RDAs) 

specific to pregnant and lactating women have been developed for a number of essential 

elements (NAS/NRC, 2000, 2001) and should be considered by risk assessors looking at these 

metals. Additionally, women lose significant amounts of Fe during menstruation, which may 

lead to increased absorption and toxicity of Cd (Berglund et al., 1994). 

4.2.3.4. Concurrent Damage or Disease 

In general, people with allergies and those with pre-existing medical conditions have 

higher-than-average biological sensitivity to environmental stressors. For example, diseases or 

treatments that damage the kidney or liver may increase sensitivity to metals that target these 

organs. Damage to the lung from smoking can potentiate effects of simultaneously or 

4-6
subsequently inhaled metals, particularly those that act directly on the lung (e.g., Be, Cd, Cr, and 

Ni). Skin abrasions or other irritations also can alter exposures to and subsequent effects of 

metals (although dermal absorption is not a primary route of metals exposure for intact skin). 

4.2.3.5. Nutritional State 

Risk assessors should be aware that dietary differences can affect the absorption of 

metals, thus changing internal target dose. For example, lack of protein (or the type of protein) 

can decrease metal uptake, thus reducing potential risk. However, relatively little is known 

about this risk factor and nutritional state is an area for further study. 

4.2.3.6. Genetic Polymorphisms and Variability 

Risk assessors should be aware of several well-known, heritable, genetic polymorphisms 

that affect susceptibility to metals. The best known of these are two disorders that affect Cu 

metabolism: Wilson’s disease and Menkes syndrome. Wilson’s disease is an autosomal 

recessive abnormality (prevalence of 1 in 30,000) that causes impaired biliary excretion of Cu, 

resulting in accumulation in and damage to various tissues, particularly the liver, brain, kidney, 

and cornea; hemolytic anemia also can result. Menkes syndrome is an X-linked recessive 

disorder of Cu metabolism (prevalence of 1 in 200,000) that resembles Cu deficiency regardless 

of level of Cu intake (WHO/IPCS, 2002). 

Hemochromatosis is another common inherited disorder. It is characterized by excessive 

Fe absorption, elevated plasma Fe concentration, and altered distribution of Fe stores (altered 

iron kinetics). One long-term effect is liver cirrhosis, with increased risk of liver cancer 

(NAS/IOM, 2003). Another Fe-related genetic polymorphism affecting Pb metabolizing 

enzymes is delta-aminolevulinic acid dehydratase (ALAD), which has been found to affect the 

relationship of bone Pb to the cumulative blood index, suggesting that the transfer of Pb from 

blood to bone is altered. It is suspected that genetic polymorphisms also exist for As metabolism 

(NAS/NRC, 2001), but these have not yet been defined. 

Risk assessors should consider the possibility of genetic differences when assessing 

potential sensitization reactions. Chronic beryllium disease (CBD) is an immune response, with 

sensitivity determined by major histocompatibility (MHC) class II genes (U.S. EPA, 1998d). 

Similarly, sensitivity to Ni is related to the genotype of human leukocyte antigens (HLA)3 

(ATSDR, 2003). 

3

 The major histocompatibility complex is a group of genes on chromosome 6 that code for the antigens that 

determine tissue and blood compatibility. In humans, histocompatibility antigens are called human leukocyte 

antigens because they were originally discovered in large numbers on lymphocytes. There are thousands of 

combinations of HLA antigens. 

4-7
4.2.4. Environmental Release, Transport, and Fate 

The exposure component of a human health risk assessment includes source 

characterization and analysis of fate and transport of the chemical of interest through 

environmental media. Models for transport and fate of metals from emission points to 

environmental receptors of importance to humans (e.g., soils, plants, or animals used in food and 

fiber) are covered elsewhere in this Framework document (see Section 3.2), as they are similar 

for both human health and ecological risk assessments. It is recommended the risk assessor 

conduct this portion of the assessment simultaneously for both human health and ecological 

assessments to ensure appropriate integration of the results. Human activities that affect the 

contact time of people with contaminated media also influence the route(s) and total amount of 

exposure. 

4.2.5. Route-Specific Differences in Effects 

Risk assessors should consider how route of exposure affects metal bioavailability and 

whether effects will occur at portal-of-entry or will be due to systemic, target-organ responses. 

Interactions among metals or other exogenous or endogenous compounds also can affect 

bioaccessibility of metals and are route dependent. Thus, many metal exposure issues are 

specific to the route of entry and will be discussed separately in the following sections. 

4.2.5.1. Inhalation Exposure 

Most airborne metals, with a few important exceptions (e.g., Hg and arsine) occur in 

particulate form. This necessitates certain considerations for inhalation exposure assessment, 

e.g., how particle size affects respirability (i.e., how much of the pollutant enters the respiratory 

system). Additionally, inhalation dosimetry for particles involves some distinctly different 

processes than for gases (i.e., deposition, clearance, dissolution, etc.), which are also influenced 

by particle size (U.S. EPA, 2004, 1997c). Particle size is thus an important factor in assessing 

metals exposure, with the focus generally being on particles less than or equal to 10 microns (F) 

in diameter (PM10). Larger particles usually do not penetrate far into the respiratory tract and 

can be cleared to the ingestion route and swallowed. Larger particles may have a larger role as 

an irritant, affecting a person’s eyes and nasal passages, and, if deposited in the uppermost 

reaches of the respiratory tract, may be transferred to the digestion tract. Thus, for exposure 

assessments involving measurements (e.g., using area or personal samples), the particle size is an 

important factor in determining inhalation exposure to metals. 

Since inhalation is a primary route of exposure for metals, the risk assessor should have a 

good understanding of inhalation dosimetry methods and how inputs vary for metals. Key 

methods for inhalation dosimetry are described in EPA guidance documents (U.S. EPA, 2004, 

1997c) and a number of models are available for calculating relative regional respiratory tract 

4-8
deposition in rodents and humans (reviewed in U.S. EPA, 2004). The guidelines for reference 

concentration (RfC) development (U.S. EPA, 1990) cite the regional deposited dose ratio 

(RDDR) model, which has been used for development of a number of RfCs for metals. The 

multipathway particle dosimetry model (MPPD) developed by the Chemical Industry Institute 

for Toxicology (CIIT) and the Dutch National Institute of Public Health and the Environment 

(RIVM) was used in EPA’s Air Quality Criteria for Particulate Matter (U.S. EPA, 2004). 

MPPD improves lung dose estimations by considering life-stage-specific parameters, particle 

clearance from the lung, and differences in oronasal breathing patterns with work load. The 

human equivalent concentration (HEC) is the concentration that is believed to result in the same 

dose to the respiratory tract region of interest as was received by the experimental animal 

species. 

In developing inhalation exposure estimates, the risk assessor should pay careful attention 

to the form of the metal pertinent to the dose-response assessment (e.g., RfC, IUR). Simply 

measuring the total amount of a metal without regard to speciation may introduce uncertainties 

into inhalation exposure estimates or other exposure routes. Metal speciation affects a range of 

processes that change how the metal is deposited in the respiratory tract and subsequently 

distributed throughout the body and, consequently, its potential toxicity (Bailey and Roy, 1994; 

Oberdorster, 1992). For example, in assessing the risk of inhaled Cr, the assessor should 

consider speciation (e.g., Cr+3 vs. Cr+6), as the dose-response assessment includes that 

specification. The bioavailability of metals via inhalation can be much higher than that of other 

intake routes. This may result in relatively high internal doses, even when intakes are similar to 

those from other routes. An example is the large contribution made by cigarette smoking to the 

body burden of Cd (e.g., Friis et al., 1998; Ellis et al., 1979). Variations in airway structure and 

respiratory conditions (e.g., as with age) may alter the deposition pattern of inhaled particles and 

contribute to variations in bioavailability (James, 1994; Xu and Yu, 1986; Phalen et al., 1985). 

For more information on the consideration of particle size in the dose-response assessment for 

RfCs and IURs, the risk assessor should refer to U.S. EPA, 1990. For metals having alternative 

Agency-developed dose-response metrics (e.g., blood Pb concentration), respirability, 

deposition, and clearance as well as absorption into the circulatory system may need to be 

addressed as part of the Exposure Assessment. 

Risk assessors should be aware of emerging issues in inhalation dosimetry that may have 

important impacts on exposure assessments for metals (U.S. EPA, 2005). The developing 

literature suggests that current dosimetry models and traditional dose measures (such as 

concentration in mass/unit volume) may not adequately characterize human health risk to very 

small particles, such as particles <2.5 μm (PM2.5) or <1 μm in diameter (nanoparticles). Much of 

the recent work on nanoparticle deposition has been conducted with metal oxide particles (e.g., 

titanium dioxide), and a growing body of literature is becoming available to the risk assessor. 

The bioavailability of Pb and other metals appears to increase with decreasing particle size, 

4-9
particularly from the inhalation and oral routes of exposure, so risk assessors should be aware of 

the potential implications for bioavailability of metal nanoparticles. 

Risk assessors should also consider exposure to metals in shower water, in which 

aerosolization can occur from the hot water tap. Though the magnitude of exposure from 

showering is unknown and comparable models do not exist for aerosolized metals (Wilkes, 

1998), models have been developed to predict human inhalation exposures due to volatile 

organics from showering (e.g., Guo, 2002; Moya et al., 1999; McKone, 1987). Where 

appropriate, the risk assessor should address such inhalation exposures during the Risk 

Characterization phase. 

4.2.5.2. Dietary Exposure 

Risk assessors should be aware that dietary 

pathways represent a major exposure route for Dietary Exposure 

metals (Choudhury et al., 2001). Estimation of Due to the diversity of the human diet, 

there may be wide variability in the intake 

intakes of metals occurring in food requires of some metals in food (e.g., seafood), 

resulting in both temporal variability (e.g., information on the levels of metals in food and the spikes after a seafood meal) and 

amount of food consumed (NAS/IOM, 2003). A geographic or cultural variability. 

number of references provide assessors with 

national-scale information on dietary exposure to metals (Capar and Cunningham, 2000; Schoof 

et al., 1999a, b; Thomas et al., 1999; Bolger et al., 1996; Dabeka and McKenzie, 1995; 

Gunderson 1995; Tsuda et al., 1995; Dabeka et al., 1993). Although large-scale surveys of the 

metal contents of foods and food consumption patterns have been conducted (e.g., Egan et al., 

2002; Ryan et al., 2001; U.S. FDA, 2001; O’Rourke et al., 1999; Thomas et al., 1999; U.S. 

DHHS, 1996), assessors should be aware that these surveys have several limitations for 

applications to human health risk assessment. Analysis is often conducted with “market basket” 

samples of packaged processed foods. With a few exceptions, such applications have not been 

empirically evaluated against biomarkers of exposure (Clayton et al., 2002, 1999; Choudhury et 

al., 2001). Risk assessors should be mindful that food consumption surveys are generally limited 

to short-term consumption (e.g., 1-3 days) and do not capture intra-individual variability that 

would affect long-term averages. Furthermore, dietary patterns may change over time (e.g., 

consumption of ethnic foods in childhood may change later), and, thus, patterns discerned at any 

given time may not accurately represent historical or future exposures. An additional challenge 

facing the risk assessor is integration of data from separate metal residue and food consumption 

surveys (e.g., Tomerlin et al., 1997). This leads to considerable uncertainty in estimates of metal 

exposure via the dietary route. 

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4.2.5.3. Incidental Soil Exposure 

Infants and children can have enhanced exposures to metals through the pathway of 

surface dust because (1) they crawl and play in close proximity to surface dust and (2) they often 

mouth their hands (e.g., finger sucking) and objects in their environment. This causes an intake 

of surface dust that is generally greater than that which is normally found in adults (e.g., Barnes, 

1990). On the other hand, infants have a large salivary response (i.e., they drool and spit up 

frequently), which may act to reduce overall dust intake. However, risk assessors should be 

aware that data are limited with regard to distinguishing between the quantity of dust ingested 

and the quantity of soil ingested. This parameter is important in connecting measured soil metal 

concentrations with surface dust ingestion that occurs in the indoor and outdoor environments 

(U.S. EPA, 1994a). Exposure assessment methods for direct soil ingestion are described in the 

Risk Assessment Guidance for Superfund (RAGS) 

(http://www.epa.gov/superfund/programs/risk/ragsa/index.htm). Additional guidance with 

respect to children (e.g., amount of soil a child may ingest) can be found in the Child-Specific 

Exposure Factors Handbook (U.S. EPA, 2002b). Few studies of soil ingestion in adults have 

been conducted; however, the estimates support the assumption that average daily soil ingestion 

rates of adults who do not participate in activities in which intensive exposure to surface dust and 

soil occur (e.g., occupational gardening, construction work) are lower than those of children 

(Calabrese et al., 1990; Hawley, 1985). Because concentrations of the metal contaminants in soil 

can be expected to vary with depth, risk assessors should consider soil metal concentrations at 

the depth appropriate to the metal(s) of concern as well as human behaviors and activities. 

4.2.5.4. Drinking Water Exposure 

Treatment of surface and/or ground water for human consumption removes dissolved 

organic carbon and suspended organic sediments that can form complexes with metals (AWWA, 

1999). Thus, inorganic forms of metals in treated drinking water will often consist of the more 

bioavailable, water-soluble species. Treatment also removes bacteria that can participate in 

organification reactions of toxicological significance to humans (e.g., methylation of inorganic 

mercuric mercury). 

Risk assessors estimating the intake of metals in drinking water will require information 

about concentrations of metals in the water and the amount of water consumed. Data on the 

metal content of tap water can be obtained from EPA’s Office of Drinking Water. EPA’s 

Exposure Factors Handbook contains exposure information on daily drinking water ingestion 

and incidental ingestion of water during swimming and showering (U.S. EPA, 1997e). 

Generally, water metal concentrations are measured at the distribution point for 

municipal water delivery systems. Distribution systems within homes (pipes, storage containers, 

etc.) can contribute significant amounts of metals (e.g., Pb, Cu) to the home drinking water 

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(Graziano et al., 1996); consequently, the contribution of metals from home-based pipes, etc., is 

a source of uncertainty in the human health risk assessment. 

4.2.5.5. Dermal Exposure 

Metals absorption through the skin is limited because the dermal route of exposure is of 

less concern during a health risk assessment. However, some metals (e.g., Ni and Cr) have the 

potential to induce toxic and sensitization effects directly on the skin (U.S. EPA, 2001, 1992). 

Dermal exposure can also lead to intakes via other routes, such as oral exposure via hand-tomouth transfer or ocular contact. 

Potential sources of dermal uptake that the risk assessor should consider include small 

particles in contact with the skin; metal exposure during bathing, showering, and swimming 

(NAS/NRC, 2002); and the uptake of metals through damaged skin (e.g., irritated skin, sunburn). 

Dermal contact with metals in soil also represents a potential route of exposure, but the relatively 

low lipid solubility of most metals limits absorption through the skin (Paustenbach, 2000; 

Hostynek et al., 1998). Few studies have actually attempted to quantify the extent or kinetics of 

the dermal penetration of metals deposited on the skin, and the applicability of these studies to 

metal species and complexes that occur in surface dust or soil is highly uncertain. 

4.2.6. Integrated Exposure 

Approaches to integrating exposure across pathways and physiological routes of uptake 

include modeling, estimates of relative bioavailability, and the use of biomarkers. 

4.2.6.1. Modeling 

Risk assessors have access to only a few specific, integrated exposure models for metals. 

The Integrated Exposure Uptake Biokinetic (IEUBK) model for Pb in children (U.S. EPA, 

1994a; White et al., 1998) was specifically developed for translating exposure measurements into 

risk estimates. The IEUBK model and background documentation are available on line at: 

http://www.epa.gov/superfund/lead/products.htm. 

Risk assessors should not apply this model to other metals because it was derived using 

Pb-specific information and consequently is pertinent only to Pb. A multipathway exposure 

model specific for As has also been developed (Cohen et al., 1998). Less complex models 

linking adult exposures and blood Pb concentrations are available (Carlisle, 2000; Stern, 1996, 

1994; U.S. EPA, 1996b; Bowers et al., 1994; Carlisle and Wade, 1992), and a stochastic human 

exposure model for Pb that is linked to a lead pharmacokinetics model may also be of use to the 

risk assessor (Beck et al., 2001; O’Flaherty, 1995). 

Other models available to risk assessors are EPA’s Total Risk Assessment Methodology 

(TRIM), which is being developed for multipathway risk assessment for air pollutants including 

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metals (http://www.epa.gov/ttn/fera/trim_gen.html); EPA’s Stochastic Human Exposure and 

Dose Simulation (SHEDS) model, a probabilistic, physiologically based model that simulates 

aggregate human exposures and doses for various population cohorts of interest (Dang et al., 

2003; Zartarian et al., 2000); EPA’s Dietary Exposure Potential Model (DEPM) (Tomerline et 

al., 1997); RESRAD, a generic exposure model developed by the U.S. Department of Energy for 

risk assessment of radionuclides (U.S. DOE, 2001; LePoire et al., 2000); and 3MRA, a 

multimedia, multi-pathway, multi-receptor exposure model developed for screening-level riskbased assessment of chronic exposures to chemicals released from land-based hazardous waste 

management units (http://www.epa.gov/ceampubl/mmedia/3mra/). 

4.2.7. Biomarkers 

Risk assessors may find biomarkers of 

exposure, effect, and susceptibility useful as basic 

tools (IPCS, 2001, 1994). Centers for Disease 

Control (CDC) and the National Health and Nutrition 

Examination Survey (NHANES) are currently 

developing a national database to quantify and 

characterize body burdens (based on human blood 

and urine surveys) that includes Pb, Hg, Co, uranium 

(U), antimony (Sb), barium (Ba), Be, Cs Mo 

platinum (Pt), thallium (Tl), and tungsten (W) (CDC, 

2005). Risk assessors can use this survey as a 

baseline measure against which the levels in receptor population individuals can be compared. 

The data are summarized in age, gender, and ethnicity categories. NAS also has completed a 

substantial amount of work in this area (NAS/NRC, 2006). 

Integration of exposures across media, route, and time of exposure can be reflected in 

biomarkers of exposure. A biomarker of exposure is a measure of cumulative exposure to a 

metal and also of metal actually existing in the body, as occurs with chronic exposure to metals. 

However, such an approach may not be appropriate for metals that are not extensively 

bioaccumulated in tissues, and it does not differentiate between metal present in a tissue in a 

sequestered or inactive form and metal engaged in toxic or pathological processes. The approach 

also does not differentiate naturally occurring exposures from those due to added metals. For 

example, arsenobetaine is a nontoxic organic form of As found naturally in shrimp and other 

seafood. The analysis of total, unspeciated urinary As, without recognition of an individual’s 

dietary history, could lead to an overestimation of exposure if the risk assessor does not account 

for seafood consumption (NAS/NRC, 1999). Thus, use of biomarkers increases the need for 

comprehensive, multipathway assessments of exposure. When available, reference or baseline 

levels of biomarkers of exposure should be incorporated into the assessment. 

Source: Henderson et al. (1989); IPCS-EHC #155 (1994). 

Hypothetical Relationships 

Among Biomarkers of Exposure 

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4.3. HAZARD CHARACTERIZATION 

Hazard identification (or hazard characterization) is “the determination of whether 

exposure to an agent can cause an increased incidence of an adverse health effect, such as cancer 

or birth defects, and characterization of the nature and strength of the evidence of causation” 

(NAS/NRC, 1994b). This includes identification of the target organ(s), consideration of any 

route-specific issues, evaluation of the adversity of the effects observed, and consideration of 

relevance to humans. Key points and metals-specific concepts to be considered in hazard 

characterization are detailed in the following sections. 

4.3.1. Mixtures and Interactions 

In most settings, individual metals exist as components of mixtures with other metals 

and/or organic substances (ATSDR, 2004; NRC, 1988). Effects of the metals in mixtures may 

be synergistic, additive, subadditive, potentiating, and/or antagonistic. Interactions among 

metals occur by competition for binding locations on specific enzymes or on cellular receptors 

during the processes of absorption, excretion, or sequestration at the target site. The presence, 

amounts, and interactions of all inorganic metals are important when considering and evaluating 

the effects of exposure and resulting human health risk assessment. 

Metals usually exist as components of mixtures with other metals and/or organic 

substances (ATSDR, 2004; NRC, 1988). Because the information or guidance on risks of metal 

mixtures is limited, risk assessors should follow published guidance for the human health risk 

assessment of chemical mixtures in general (U.S. EPA, 2000b, 1986) and cumulative risk 

guidance (U.S. EPA, 2003e). 

Only a few controlled studies exist on the interactions of metals relevant to levels found 

in the environment (see ATSDR, 2004, on mixtures of (1)As, Cd, Cr, and Pb; and (2)Cu, Pb, Mn, 

and Zn). Risk assessors may use the current default approach, which assumes additivity of the 

doses for each metal as it will produce estimates that are overly-cautious. This approach, which 

involves calculation of a hazard index (HI), is 

most appropriate for chemicals that produce the 

same effects by similar modes of action Metal-Binding Proteins 

Metallothioneins (Ag, Hg, Cu, Bi, Cd, Pb, Zn) t (MOAs). However, differing potencies of Transferrin (Fe, Al, Mn) t 

metals with similar MOAs should be accounted Ferritin (Fe, Cd, Zn, Al, Be) s 

Ceruloplasmin (Cu, Fe) tr

for by converting chemical concentrations into Lead-binding protein(s) (Pb) s 

an equitoxic dose using either toxic units (TUs) Membrane carrier proteins t 

or toxicity equivalence factors (TEFs). In the s = storage; t = transport; tr = transform 

case of chemicals with different MOAs, the risk 

assessor should consider estimating separate effect-specific HIs for each chemical in the mixture 

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using the RfD as the toxicity value for each effect. Newer EPA guidance provides a number of 

quantitative approaches for characterizing mixture risks (described in detail in U.S. EPA, 2000b, 

1986). 

The terms molecular mimicry and ionic mimicry have been applied to situations in which 

a metal forms a complex with an endogenous ligand and the resulting compound mimics the 

behavior of a normal substrate, disrupting normal function (Ballatori, 2002; Clarkson, 1993). 

Molecular or ionic mimicry may be viewed as a form of metal-metal interaction. Most examples 

involve the replacement of an essential metal with a nonessential metal. For example, Cd can 

mimic and substitute for Zn and Ca. Additionally, many different proteins in the body complex 

with metals which may modify their toxicity and kinetics (e.g., some metals bind with albumin 

for purposes of transport in the circulatory system and across cell membranes or within cells). 

Some proteins have different binding kinetics for the various metals, resulting in specific proteinmetal interactions. Risk assessors should be familiar with these metal-binding proteins to 

correctly interpret the bioavailability of individual metals within a mixture and the potential use 

of protein expression as a biomarker of metal exposures. 

Many of the interactions between essential metals are related to maintaining optimal 

nutritional levels by synergisms and antagonisms at both physiological and extrinsic (dietary) 

sites. World Health Organization (WHO) publications (IPCS, 2002; WHO, 1996) have 

summarized these homeostatic interactions, which are often complex (e.g., excess Ca in the diet 

may induce signs of Zn deficiency, even if the Zn intake is normal). Similarly, excess Zn in the 

diet may aggravate Fe deficiency. 

Interactions between essential and nonessential metals are very common (e.g., Cd uptake 

can mimic that of Zn). Similarly, among anions, mimicry of the sulfate and phosphate ions 

occurs. However, the risk assessor should be aware that validated data in humans are rare, and 

that applications of this phenomenon are best limited to screening-level assessments. Two 

nonessential metals may compete for passive transport across common sites on a membrane or 

with an essential metal on an active binding site. These effects may not be additive and likely 

relate to relative binding strength. One metal may affect one site and another metal may affect a 

different site; this can include both active and passive transport or binding sites and may or may 

not include interactions with essential metals. These effects often will be additive. The risk 

assessor should be familiar with the MOAs of each metal of concern to develop at least a general 

understanding of whether the mixture effect is likely to be additive or more than additive. 

Metals can be active at most cellular sites where organic toxicants have their effects and 

directly interfere with receptor activation (Stoica et al., 2000), ion channel regulation (Kiss and 

Osipenko, 1994), cell signaling (DeMoor and Koropatnick, 2000), cell adhesion (Prozialeck et 

al., 2002), and gene transcription (Meplan et al., 2000). Thus, metals are not readily 

distinguished from organic substances in the range of their potential mechanisms of action at the 

cellular and molecular level, so the risk assessor can transfer knowledge about toxicity across 

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chemicals. Additionally, co-occurrence of metals with organic substances can change the 

bioavailability and increase or decrease absorption. For example, in the diet, citrates and 

histidine are known to enhance Zn absorption, whereas ascorbate can modify Fe-Cu 

antagonisms. Low protein content may increase the absorption of Cd and Pb, and oral 

contraceptives may influence the metabolism of Fe, Cu, and Zn. The risk assessor should be 

aware of these conditions to avoid over- and underestimating risk. 

4.3.2. Essentiality Essentiality

Certain elements are nutritionally essential to Essentiality should be viewed as 

humans and play a key role in physiological or part of the overall dose-response 

biochemical processes (NAS/IOM, 2003; IPCS, 2002; relationship. The shape of this 

relationship can vary among 

WHO, 1996). Elements essential to other organisms organisms. For a given subpopulation, 

reference doses designed to protect 

may not be essential to humans and vice versa. from toxicity of excess should not be 

Adverse nutritional effects can occur if essential metals set below doses identified as essential 

for that subpopulation. are not available in sufficient amounts, and nutritional 

deficits also can be adverse and increase the 

vulnerability of humans to other stressors, including those associated with other metals. 

Essentiality should be viewed as part of the overall dose-response relationship, and reference 

doses designed to protect from toxicity of excess should not be set below doses identified as 

essential. 

Metals that are currently deemed nutritionally essential for humans are Co, Cr III, Cu, Fe, 

Mg, Mn, Mo, Se and Zn (Table 4-1). Some metals (e.g., B, Ni, Si, V, and perhaps As), while not 

essential to human health, may have some beneficial effects at low levels of exposure 

(NAS/IOM, 2003). Risk assessors should consider the essential elements as comprising three 

groups: those that are cations (Zn, Fe, Cu, Mn, Cr), those that are anions (Mo, Se), and those 

that are a bio-inorganic complex (i.e., the Co complex, cobalamin). The homeostatic 

mechanisms differ for each group, and risk assessors can use this knowledge to generally classify 

the types of health effects that might occur and the potential bioavailability of the metal. In 

general, the gastrointestinal tract and the liver regulate the uptake and transfer of cations (e.g., 

Fe, Zn, Cu). The anionic group is more water-soluble and is less reactive with N, S, P, O, and 

hydroxide groups than are cations. They are absorbed very efficiently through the intestine and 

their subsequent compartmentalization and excretion is by manipulation of their oxidation and 

methylation states; total body burden is regulated by renal excretion. The risk assessor should be 

aware that homeostatic controls do not typically apply to effects at the portal of entry, so effects 

can be seen at lower doses than those required for systemic responses. 

The risk assessor should view essentiality as part of the overall dose-response 

relationship. The risk assessor should use the entire dose-response relationship, from very low 

(inadequate) doses to high (toxic) doses (see text box) when determining an acceptable upper 

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Table 4-1. Metal essentiality for humans 

Nutritionally essential metals Nutritionally nonessential metals 

Cobalt II, III 

Chromium III 

Copper 0, I, II 

Iron II, III 

Magnesium II 

Manganese II, IV 

Molybdenum IV, VI 

Selenium II, IV, VI 

Zinc II 

Aluminum III 

Antimony III, V 

Arsenic III, V 

Barium II 

Beryllium II 

Bismuth III, V 

Boron III 

Cadmium II 

Cesium* I 

Chromium VI 

Gallium* III 

Germanium* IV 

Gold 0, I, III 

Indium* III 

Lead II, IV 

Lithium I 

Mercury 0, I, II 

Nickel II 

Niobium* V 

Palladium* 0, II 

Platinum* 0, II, IV 

Rubidium I 

Silicon* IV 

Silver 0, I, II 

Strontium II 

Tellurium* II, IV, VI 

Thallium I, III 

Tin II, IV 

Titanium IV 

Tungsten VI 

Uranium IV, VI 

Vanadium III, V 

Zirconium* IV 

* Limited human data for these metals. 

exposure limit. Several agencies have 

developed guidance for selecting a 

benchmark dose that is not too low (and, 

therefore, likely to result in deficiency) or 

too high (and likely to result in toxicity to 

some segment of the population). IPCS 

(2002) guidance describes the use of an 

“Acceptable Range of Oral Intake” 

(AROI), which estimates the minimal 

requirement to prevent deficiency and an 

upper limit that will produce toxicity. The NAS Food and Nutrition Board (FNB) in conjunction 

with the Institute of Medicine (IOM, 2001; NAS/IOM, 2000) developed the Dietary Reference 

Intakes (DRIs) program and reformulated RDAs (now known as RDIs) using the estimated 

average requirement (EAR) or adequate intakes (AIs). They also developed a tolerable upper 

intake level (UL), which is the highest level of daily nutrient intake that is likely to pose no risk 

of adverse health effects to almost all individuals in the general population. The UL is based on 

a risk assessment model similar to that used by EPA to set the RfDs and is intended to protect the 

population from adverse health effects resulting from excess exposure to a compound. ULs are 

available for all the essential metals and for B, Ni, and V. ULs do not take into account sensitive 

Observed Level of Intake 

4-17
or immuno-compromised populations. ULs may differ from RfDs because they are derived from 

human studies rather than animal studies and use smaller uncertainty factors. Additionally, the 

risk assessor should be cautioned that RfDs are intended to cover sensitive subpopulations, 

whereas RDAs are estimated to satisfy the nutritional needs of 97.5% of the healthy U.S. 

population. RDAs are specific to different age groups and genders, with listings for 16 different 

age-sex and six age-pregnancy combinations (NAS/IOM, 2003). 

4.3.3. Forms of Metals 

Unlike organic chemicals, metals are neither created nor destroyed by biological or 

chemical processes. However, these processes can transform metals from one valence state to 

another and can convert them between inorganic and organic forms. Information developed for 

one form of a metal may not be directly applicable to other forms. Different valence states 

(species) of the same metal affect bioaccessibility and bioavailability, and they elicit different 

responses in the human body. The particle size and environmental matrix (water, soil, air), 

within which the metal is embedded, influence exposure amount, rate, and route, particularly for 

the inhalation pathway, which can then result in different target organs and response levels. 

Therefore, the risk assessor should consider which form of the metal of interest is being assessed. 

If exposure or effects information has been developed for a different metal species, the risk 

assessor should either make appropriate adjustments or acknowledge this as a significant 

uncertainty in the Risk Characterization. For example, Cr(III) is essential in the diet, whereas 

inhaled Cr(VI) is carcinogenic. 

Risk assessors should be aware that information developed on the health effects of one 

form of a metal may not be directly applicable to other forms, particularly organometallics; other 

guidance documents should be consulted when conducting risk assessments for organometallics. 

4.3.4. Toxicokinetics/Toxicodynamics 

Homeostatic mechanisms such as binding of 

metals to proteins can introduce significant complexities 

to hazard assessments for metals, with significant 

quantitative effects once these mechanisms are 

overwhelmed. Certain metal compounds bioaccumulate 

in human tissues, and it is important to recognize that 

such bioaccumulation is related to toxicity (SAB, 2006; 

see Section 2.3.1 for a definition and more detailed 

discussion of bioaccumulation). Since not all tissues may 

be of “interest” from a human health perspective, net accumulation in human tissues may or may 

not be relevant in the hazard characterization of metals (see, for example, Section 4.3.4.2 for a 

Toxicokinetics 

Toxicokinetics describes the series 

of processes that dictate the 

disposition of a substance in or on the 

body after exposure occurs and 

processes related to deposition on 

the body surfaces (also considered in 

exposure assessments), absorption, 

distribution, metabolism, and 

excretion (ADME). 

4-18
discussion of metal sequestration and subsequent remobilization). Further, there are no available 

simple metrics that allow quantification of the potential for human bioaccumulation of metals, 

although a full pharmocokinetic model can be used to 

estimate metals bioaccumulation and distribution in human Bioaccumulation of metals is the net 

tissues. Additional discussion of the potential for accumulation of a metal in the tissue 

of interest or the whole organism that 

remobilization of sequestered metals is in section 4.3.4.2. results from all environmental 

All these processes can be described through the use of exposure media, including air, water, 

solid phases (i.e., soil, sediment), and 

physiologically based pharmacokinetic (PBPK) models. diet, and that represents a net mass 

balance between uptake and Integrated descriptions provide metrics of internal dose elimination of the metal (SAB, 2006). 

(including biological markers of exposure) that can be used 

by risk assessors to improve the quantitative basis of doseresponse assessments. Unique features of metals that result in differences in toxicokinetic 

behavior of metals as compared to organic substances are shown in Table 4-2. 

Table 4-2. Summary of major differences in kinetic behavior of organic 

compounds compared to metals and inorganic metal compounds in humans 

Organics Metals 

Tissue uptake is most commonly a blood flowlimited process, with linear portioning into tissues. Metals and their complexes are often ionized, 

with tissue uptake (membrane transport) having 

greater potential to be diffusion-limited or to use 

specialized transport processes. 

Metabolism is generally extensive and often speciesspecific. Metabolism is usually limited to oxidation state 

transitions and alkylation/dealkylation reactions. 

Persistence in body fat is common because of lipid 

solubility (not capacity-limited). 

Often sequestered, bound to specific plasma or 

tissue proteins (intrinsically capacity-limited) or 

bone. 

Due to complex metabolism, organics may be 

eliminated by excretion in urine after 

biotransformation from lipophilic forms to 

hydrophilic forms, in bile after conjugation to large 

organic molecules, or in exhaled air if not 

metabolized. 

Predominantly eliminated in urine because metal 

compounds are generally small molecules and are 

hydrophilic. As a result of protein binding, may 

be excreted via hair and fingernails. 

Generally substance-specific homeostatic 

mechanisms are not available. 

Essential metals have homeostatic mechanisms 

that maintain optimum tissue levels over a range 

of exposures. 

Interactions with other structurally similar 

compounds may occur, especially during 

metabolism. 

Interactions among metals and between metals 

and organics are numerous and occur commonly 

during the processes of absorption, excretion, and 

sequestration. 

Source: Adapted from Golub et al. (2004). 

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4.3.4.1. Absorption 

Absorption is a process by which an administered substance enters the body. 

Bioavailability is a term often used to describe the degree of absorption. Two elements of the 

absorption process are critical for evaluating systemic doses of metal compounds, both the 

degree and rate of absorption. Although information on both of these parameters is ideal for 

developing a quantitative estimation of systemic doses, information on the degree of absorption 

is more commonly available for most chemicals. Metals have a number of unique properties that 

impact their absorption across biological membranes. Key factors affecting the absorption of 

metals include solubility, particle size, valence state, lipophilicity, and the exposure matrix. 

Soluble forms of the metal are more readily absorbed since the metal ion itself is typically the 

absorbed entity. The bioavailability of metals increases as particle size decreases. However, in 

the lung particle size also determines the site of deposition and thus the clearance mechanisms 

that can ultimately result in systemic uptake (via transport to the lymph system following 

macrophage engulfment) or the GI tract (via mucociliary transport). Chemical speciation in 

terms of valence state can affect absorption. Recent progress in identifying metal transporters 

suggests that generalizations are not appropriate, and each metal should be assessed in terms of 

its ability to access transporters and the presence of transporters in potential target organs. In 

general, lipophilic compounds will be absorbed more readily than hydrophilic ones. For 

example, human skin is not very permeable, and it provides a good barrier against dermal 

absorption of metals and metal compounds; elemental Hg and dimethyl Hg are notable 

exceptions. Risk assessors should note the complexities in absorption processes since they have 

direct implications on metals risk assessment, primarily in requiring detailed consideration in 

extrapolating across different exposure conditions or animal species. Absorption can vary 

dramatically for different forms of the same metal, for the same form of metal in different 

matrices, among different species, and across different routes of exposure. Therefore, it is not 

appropriate to assume concordance in absorption of metal compounds without a detailed 

evaluation (and documentation) of the scientific basis for such an assumption. For example, 

empirical information on dermal absorption of metals should be consulted when available 

(Stauber et al., 1994; Wester et al., 1992; Hursh et al., 1989; Ilyin et al., 1975), and similar 

considerations apply to other routes of exposure. 

4.3.4.2. Distribution 

The unique features of metals influence their distribution to potential target organs and 

the subsequent target tissue doses. The distribution of metals reflects their transport and 

accumulation in the body within tissues, blood or plasma, or other extracellular space. 

Partitioning to blood and cellular components, particularly via interactions with proteins, is of 

particular importance for metal risk assessment. Retention in tissues of metals or metal 

4-20
compounds generally is related to formation of inorganic complexes or metal protein complexes 

(e.g., Pb in the bone compartment and Cd in tissues bound to low-molecular-weight 

metallothionein proteins). Risk assessors should be aware that retention of metals in the body by 

protein binding or sequestration in a nontoxic form allows the metal to reside in the body without 

producing a toxic or pathological effect. For example, As and Hg have relatively short 

biological half-lives that can be measured in days, whereas Cd and Pb can be bound or 

sequestered in inactive forms for years. Cd is retained in soft tissues (e.g., liver and kidney) for 

10 to 20 years by intracellular binding with metallothionein. Metal binding to proteins is 

capacity-limited, and toxicity to target organs occurs when the binding capacity is exceeded. 

Thus, the potential for toxicity exists in older adults for metals with long half lives that are 

initially adequately sequestered. Conversely, retention in tissues is a dynamic equilibrium and 

can be a source of internal exposure long after the external exposure source has been removed. 

For example, Pb in bone can be mobilized during pregnancy and lactation or as a result of 

osteoporosis (USEPA IRIS, http://www.epa.gov/iris/subst/0277.htm). The risk assessor should 

consider all the aspects of metal distribution in the body (i.e., binding and sequestration plus 

release processes) when estimating likely target dose. The risk assessor should also note the 

uncertainty associated with lack of information to complete a quantitative analysis of these 

processes during the Risk Characterization. 

4.3.4.3. Metabolism 

Metabolism of metals is limited to oxidation-reduction reactions or alkylation/ 

dealkylation reactions. In these reactions, new inorganic species or metal organic complexes 

may be formed, but the metal ion persists. Nevertheless, differences in these transformation 

pathways among human populations or across species have practical implications for risk 

assessment because different species of the same metals often have very different toxicities. 

Because of this, the risk assessor should fully explore available data on metal metabolism. For 

example, As can be metabolized to organic forms that are less toxic than the inorganic As to 

which an individual is initially exposed. 

4.3.4.4. Excretion 

Risk assessors should be aware of the number of qualitative differences in metal 

excretion as compared to organic compounds. These include the greater likelihood for excretion 

in urine and the propensity for metals to be excreted via hair and nails. In addition to reducing 

the target dose, risk assessors can use these excretion routes to develop biological markers of 

exposure for many metals. 

Although metals share many similar characteristics, their excretion kinetics can vary 

dramatically, primarily because of differences regarding sequestration in bone or binding to 

4-21
specific proteins. This reduces the rate of excretion (or generates a biphasic excretion curve with 

two or more distinct excretion half-lives). Risk assessors should assess each metal individually 

to incorporate excretion rates in the calculations of target dose and subsequent hazard 

assessment. 

4.3.4.5. Kinetic Modeling 

Physiologically based pharmacokinetic/pharmacodynamic (PBPK/PBPD) modeling of 

metals entails the mathematical description and modeling of their absorption, distribution, 

metabolism, and excretion (ADME). A typical PBPK model consists of multiple compartments 

representing tissues or tissue groups that are linked by blood flow. PBPD models describe the 

relationship between target tissue dose and health endpoints or target tissue effects. Combined 

use of PBPK and PBPD models provides understanding of the complex relationships between 

exposure and target organ effects. Risk assessors may find that these models are valuable risk 

assessment tools for purposes of interspecies, high-dose/low-dose, route-to-route, and exposure 

scenario extrapolations (Krishnan and Andersen, 1994). A PBPK model allows the risk assessor 

to define the relationship between external exposure and an internal measure of a biologically 

effective or toxic dose in both experimental animals and humans, thus increasing the precision of 

extrapolating effects thresholds to humans. Use of PBPK models can account for nonlinear 

uptake, metabolism, and clearance; toxicity associated with products of metabolism rather than 

the parent chemical only; and tissue interactions. The underlying assumption of PBPK is tissue 

dose equivalence, i.e., that health effects are caused by the toxic form(s) of the chemical 

measured at the biological target (Krishnan and Andersen, 1994). 

PBPK models historically have been developed and used for risk assessment mainly with 

volatile organic compounds (VOCs) (e.g., methylene chloride) (Andersen et al., 1987), but 

PBPK models have recently been applied to some metals (White et al., 1998; Clarke, 1995). 

There are currently three main Pb risk assessment PBPK models. The O’Flaherty model 

(O’Flaherty, 1995) is a PBPK model for children and adults. It includes the movement of Pb 

from exposure media (i.e., intake via ingestion or inhalation) to the lungs and gastrointestinal 

tract and subsequent exchanges between blood plasma, 

liver, kidneys, and richly and poorly perfused tissues; and Metal Toxicity 

excretion from liver and/or kidney. The Leggett model The organ or tissue in which 

metal toxicity occurs may differ from (Leggett, 1993) allows risk assessors to simulate lifetime the organ or tissue(s) in which the 

exposures and can be used to predict blood Pb metal accumulates and may be 

affected by the metal’s kinetics. 

concentrations in both children and adults. The EPA has Effects at the portal of entry do not 

performed a comparison of these adult Pb risk assessment depend on bioavailability. Both the 

exposure route and the form of the 

models metal can affect a metal’s 

(http://www.epa.gov/superfund/lead/products/adultrevie carcinogenic potential and its 

noncancer effects. 

4-22
w.pdf) . EPA developed the Integrated Exposure Uptake Biokinetic (IEUBK) model to predict 

Pb levels in children (U.S. EPA, 1994a) and recommends that it be used as the primary tool for 

Pb risk assessment at Superfund and RCRA corrective action sites (OSWER Directive, 1998; 

http://www.epa.gov/superfund/lead/products/oswer98.pdf). 

Risk assessors can review differences in kinetic behavior between metals and VOCs in 

O’Flaherty (1998). In brief, these include: (1) oral bioavailability, (2) inhalation bioavailability, 

(3) cellular uptake, (4) nutritionally essential and nonessential metal interactions, (5) proteinbinding behavior and function, (6) incorporation into bone or hair, (7) metabolism, and (8) 

excretion. Moreover, risk assessors should keep in mind that many of the processes controlling 

the disposition of metals are intrinsically capacity-limited and can result in extended residence 

times. Risk assessors should use multiple lines of evidence to understand the kinetics and, 

therefore, the hazard of metal sequestration and elimination. The major challenge faced by the 

risk assessor when using PBTK models for metals is to balance the complexity of the biology 

with the data available to parameterize the model. Estimation of many parameters from the same 

data or insufficient data (over-parameterization) leads to greater uncertainty in model predictions 

and limits the utility of the model for regulatory purposes. 

4.3.5. Metal Toxicity 

Diversity in observed toxicities of different metals likely reflects the variety of 

biochemical mechanisms by which they exert their effects and variability in their toxicokinetic 

properties. At least five metals are known carcinogens, and several other effects of metals are 

also well documented, including effects on the neurological, cardiovascular, hematological, 

gastrointestinal, musculoskeletal, immunological, and epidermal systems. 

4.3.5.1. Noncancer Effects of Metals 

Metals and metal compounds have very diverse toxicological profiles. For risk 

assessment purposes, selected critical effects serve as the basis for deriving threshold or 

benchmark toxicity values (e.g., RfDs) and are defined as “the first adverse effect, or its known 

precursor that occurs to the most sensitive species as the dose rate of an agent increases” (U.S. 

EPA, 2005). Both the mechanism of toxicity and the critical effect may vary with the form of 

the metal. Additionally, short-term exposures may produce target organ effects very different 

from those produced by a similar dose over a longer period of time. Short-term, high-level 

exposure by ingestion may give rise to well-recognized acute toxicity syndromes, usually 

involving the gastrointestinal tract initially and possibly, secondarily involving the renal, 

cardiovascular, nervous, or hematopoetic systems. Survivors of acute high-dose As ingestion 

usually experience multiple organ effects, sometimes with long-term sequelae. Long-term, lowdose exposures from ingestion of metals in food and water generally cause an accumulation in 

4-23
target organs over time. Such exposures can involve any organ system but do not usually 

produce overt gastrointestinal symptoms. For example, low-level, long-term exposure to Cd in 

food—sometimes combined with inhalation exposure from cigarette smoking—will cause Cd to 

accumulate in target organs (e.g., kidney) but will not produce any obvious clinical effects until 

“excess” capacity is diminished to a point where the normal function is lost (e.g., onset of renal 

disease and/or osteoporosis later in life). 

In addition to considering systemic effects, the risk assessor should also examine portalof-entry effects. Unlike systemic effects, which may be route-independent, portal-of-entry 

effects are not observed following other routes of exposure. For example, dermal irritation, 

sensitization, and allergic responses from metals can occur without absorption and systemic 

responses. 

4.3.5.2. Carcinogenic Effects of Metals 

At least five transition metals—As, Cd, Cr(VI), Be, and Nil—are accepted as human 

carcinogens in one form or another or in particular routes of exposure (IARC, 2004b; NTP, 

2002) and inorganic Pb compounds are considered probable human carcinogens by EPA’s IRIS 

program, while IARC (2004a) has concluded that there is limited evidence of carcinogenicity to 

humans (see: http://www.epa.gov/iris/subst/0277.htm#carc and 

http://monographs.iarc.fr/htdocs/announcements/vol87.htm). Other metals have mixed evidence 

regarding potential carcinogenicity. Therefore, risk assessors should pay careful attention to 

approaches for cancer risk assessment as applied to metals. Several guidance documents are 

available for use by the risk assessor in developing or interpreting cancer risk assessments (e.g., 

U.S. EPA, 2005), as are international efforts that provide guidance on assessing human relevance 

of tumors identified in animals (Cohen et al., 2004). 

Nickel and Pb compounds and Cr and Cr compounds are well-established contact 

allergens. Other metals that have been cited as contact allergens include Cu (WHO/IPCS, 1998), 

Co salts (AIHA, 2003), organomercurials (AIHA, 2003), Be (IPCS, 1990b), palladium (Pd) 

(Kimber and Basketter, 1996), and gold (Au) (Kimber and Basketter, 1996). Although there is 

some connection between skin and respiratory sensitization, it does not follow exact rules, and 

the dermal mode is a much more common reaction to metals. 

A key consideration for the risk assessor in cancer risk assessment is the determination of 

the MOA of carcinogenesis, a general description of how the chemical causes cancer. The MOA 

determines human relevance of observed animal tumors, any route-specific differences (e.g., 

carcinogenic at the portal of entry via the inhalation route, but not carcinogenic via the oral 

route), and the approach used for extrapolation from experimental doses in animals to 

environmentally relevant human doses. In particular, the MOA evaluation is a key consideration 

in whether a linear or nonlinear approach is used to extrapolate to low doses. The MOA is 

4-24
known for some, but not all, metals. For those for which the MOA is unknown, the risk assessor 

should refer to the Cancer Guidelines (U.S. EPA, 2005) for guidance. 

4.3.5.3. Issues Related to Evaluation of Toxicity Tests for Metals 

As with any hazard assessment, risk assessors prefer a robust dataset on toxic responses 

of the metal(s) of concern for key endpoints (e.g., irritation and sensitization, systemic noncancer 

toxicity, and genotoxicity or tumorigenicity). In many cases, metals will be well-studied, and 

thus, human studies (epidemiology, controlled clinical studies, or case reports) will be available 

to aid in hazard characterization. For metal compounds for which adequate human data are not 

available, the risk assessor must rely on animal toxicity studies. U.S. EPA has established 

guidelines for assessing the adequacy of a database for derivation of chronic human health risk 

values such as RfDs and RfCs (U.S. EPA, 2002a) and has provided guidance for evaluating the 

“weight of evidence” for carcinogenicity (U.S. EPA, 2005). These generic guidelines are 

applicable to metals as well as organic compounds as long as the risk assessor considers the 

following metal-unique aspects of hazard determination: 

• Adequate controls. When a salt of a metal is administered to the test animals, the 

risk assessor should evaluate that a suitable control group was used, specifically that 

any potential for salt-induced toxic responses is appropriately assigned probable 

cause. 

• Dosing solubility, ionization, hydration, and speciation of metals administered in 

water. Metal compounds may be in suspension or in solution and may be 

differentially hydrated depending on the concentration in which they are prepared and 

the length of time the preparation stands, potentially resulting in different 

pharmacokinetic and toxic properties. Water pH and mineral content also are 

relevant factors to be considered by the risk assessor. 

• Trace element content of food and drinking water. Because of the well-known 

interaction of metals with essential trace elements, the trace element content of 

animal feed and drinking water or of vehicles used for gavage or injection studies 

should be reported or controlled. Inconsistent results across experiments could be 

due to this factor. 

• Acute stress in the experiment. A component of acute stress in the experiment can 

induce hepatic metal-binding proteins (acute phase proteins) and alter the toxicity of a 

given administered dose. 

To achieve an adequate internal dose for the study of toxicity, animal toxicologists often 

use bioavailable forms of metals. For the initial characterization of a toxicity syndrome, it is not 

practical to simultaneously test all forms of a metal that may be involved in human exposures. 

For example, Al researchers commonly use aluminum lactate, which is known to reliably 

provide elevated tissue concentrations in laboratory animals, or aluminum maltolate, which 

4-25
provides a stable ion pool in water solution. However, a risk assessor is very unlikely to conduct 

an assessment of Al in its lactate or maltolate form. Thus, the risk assessor should be aware that 

failure to adjust the toxicity data generated from water-soluble metal species to the appropriate, 

less soluble species of concern introduces uncertainty. There are both direct and indirect 

approaches to address the relative bioavailability of metals in the environment: (1) conduct new 

animal toxicology studies using the metal form encountered in the site assessment; (2) use 

adjunct scientific data to derive an adjustment to the effective dose identified in the animal study 

(e.g., data on the distribution of chemical forms of the metal in the environment or at a 

contaminated site); or (3) use a default assumption that the metal in the environmental samples is 

the same as that tested. The first approach is more scientifically sound but often is precluded by 

time and financial resource limitations; the third option generally is the most health-conservative. 

A fourth alternative to conducting site-specific assessments is for the risk assessor to 

estimate bioavailability through solubility studies or limited bioavailability studies of specific 

samples from the site. For example, arsenic bioavailability has been estimated for soils from 

various contaminated sites (Ng et al., 1998; Freeman et al., 1995, 1993) and also through a series 

of solubility studies of soil from a site contaminated with mine tailings (Ng et al., 1998; Salocks 

et al., 1996). 

4.3.6. Dose-Response Assessment 

The result of the hazard characterization is a determination of the key noncancer and 

cancer endpoints related to exposure to the metal of interest. The risk assessor then uses these 

data as the input to the dose-response assessment to “characterize the relationship between 

exposure or dose and the incidence and severity of the adverse health effect” (NAS/NRC, 

1994b). Assessors should consider the factors that influence dose-response relationships, such as 

intensity and pattern of exposure, age, and lifestyle variables. Traditionally, separate approaches 

have been used for dose-response assessment for noncancer and cancer endpoints. For 

noncancer endpoints, the result of the dose-response assessment is an RfD for oral or dermal 

exposure or an RfC for inhalation exposure. For cancer assessment, the approach depends on the 

chemical’s MOA. Classically, the result of the cancer assessment is a measurement of the risk 

per unit dose, either as a slope factor or unit risk. 

A key consideration for the risk assessor when deriving metal dose responses is to 

express the exposure potential and the toxic response as the same metal species. In general, for 

systemic effects of soluble metal salts, the risk assessor should express toxicity in terms of the 

dose of the metal ion rather than the metal salt. In contrast, if the toxicity is related to a specific 

compound, particularly for portal-of-entry effects, risk assessors should express it in terms of the 

compound, rather than the ion. These differences should be considered when applying toxicity 

values in risk assessments. For example, in IRIS, Se is listed as “selenium and compounds,” and 

there are separate assessments for “nickel, soluble salts,” “nickel subsulfide,” “nickel refinery 

4-26
dust” (a mixture), and “nickel carbonyl” (a highly reactive compound that behaves differently 

from other nickel compounds). In some situations, the toxicity of the anion needs to be 

considered. For example, there are separate IRIS documents for Ag and silver cyanide. 

However, careful consideration of the chemical form has not been applied historically to all such 

documents, and the risk assessor should carefully consider the applicability of the chosen toxicity 

values to the chemical forms of interest, paying 

close attention to solubility, bioavailability, and Considerations in Risk Characterization 

physical/chemical properties as well as Variability. Inter-individual biological 

available toxicity data. differences exist within an animal or human 

population, or measurement differences exist 

A related issue is whether the toxic owing to method imprecision. 

response is reported in terms of the added metal Uncertainty. Data are unavailable. 

Incertitude. Knowledge about key 

in the diet or the total metal (i.e., whether the relationships is not available. 

amount of trace elements in the control diet in 

the animal studies is included in the dose calculations). Risk assessors should carefully review 

the supporting documentation for toxicity values to determine precisely what is being reported 

and to account for any potential interactions among dietary metals. The assessor also should 

consider whether the study provided adequate levels of trace elements so that the observed 

toxicity is not secondary to some unrelated deficiency. 

Dose-response assessments for some metals are based on data from human occupational 

studies. While derivation of the dose-response value (e.g., the RfC, RfD, or cancer slope factor) 

will have included some steps to extrapolate the occupational study data to environmental 

exposures (e.g., dosimetric and duration adjustments), the form of the chemical may merit 

consideration. For example, the occupational exposure will have involved a particular range of 

particle sizes, which may influence the response observed (U.S. EPA, 2002a). 

Risk assessors should review RfDs derived for essential elements to ensure that they are 

not below required daily intakes. The RfD should not be below the general population RDA. 

This means that the risk assessor should give careful consideration to the appropriate size of 

uncertainty factors, which is often made easier by the frequent (although not uniform) 

availability of human data for relatively large and diverse populations. The only exception to the 

comparison between the RfD and nutritional requirements is that certain populations (e.g., 

pregnant women) may have higher nutritional requirements, while these levels could 

theoretically be toxic to other populations. In such cases, the risk assessor should be careful to 

avoid logical inconsistencies and to identify the sensitive population on which the RfD is based. 

4.4. RISK CHARACTERIZATION 

Risk Characterization is the final phase of the risk assessment process. It is the phase in 

which information from hazard characterization, dose-response assessment, and exposure 

assessment is considered together to determine the actual likelihood of risk to exposed 

4-27
Risk Characterization 

Have the qualitative assessment, quantitative 

assessment, and key uncertainties regarding 

metals been presented in accordance with EPA 

guidelines? 

Do conclusions fully reflect risks in relation to 

ambient concentrations, essentiality of metals, 

chemical speciation, and information on human 

variability in sensitivity? 

Have assumptions and uncertainties been 

documented adequately? 

Have available data on mechanisms of action 

and metal interactions been fully explored in 

developing the quantitative assessment? 

populations (NAS/NRC, 1994b). For exampl

inorganic As occurs naturally in food and 

water; thus, a risk characterization would 

integrate the currently accepted dose-response

information for inorganic As with the exposur

assessment information for a particular food o

drinking water source, or for the national 

distribution of intake from food and water, to 

determine whether a potential problem exists. 

During Risk Characterization the uncertainties

in the dose-response assessment, the 

uncertainties in the estimate of exposure 

derived for the scenario under evaluation, and the level of confidence in the overall 

determination of risk should be laid out. At the same time, Risk Characterization is the first 

phase in the risk management process, in which information from the characterization is 

integrated into the consequences of rule-making or risk management, such as consideration of 

cost, alternative solutions, political considerations, and community interactions. 

e, 

e 

r 

Risk assessors should refer to guidance on Risk Characterization (U.S. EPA, 2000c) that 

identifies key goals and steps for a Risk Characterization. Each Risk Characterization should 

include three components: a qualitative summary of each section of the risk assessment, a 

numerical risk estimate, and a description of uncertainties. Since metal exposures often occur in 

the context of mixtures (either mixtures of metals of the same form, mixtures of different metal 

elements, or mixtures with organics), risk assessors should consult additional Risk 

Characterization tools developed for mixtures (U.S. EPA, 2000b, 1986). These guidelines 

specify that the characterization of risks from mixtures of metals (and other compounds) be 

based primarily on information about the types of interactions that might be present. 

Risk assessors should include a discussion in the Risk Characterization of the sources of 

variability and uncertainty in the risk assessment process. This is particularly important for 

metals risk assessments given all the components above. These are in addition to the variability 

and uncertainties that are inherent in all risk assessments (e.g., animal-to-human toxicity 

extrapolations). Because information, knowledge, and tools are lacking for many of the metalspecific uncertainties, risk assessors should be particularly diligent in documenting whether these 

may result in an over- or underestimation of risk (i.e., result in a conservative risk estimate or 

not). It is likely that site-specific risk assessments will have fewer uncertainties than regional- or 

national-scale assessments because risk assessors have access to local data on key issues such as 

specific metal species, relative bioavailability, or current metal levels. For national or regional 

assessments, selection of ranges or specific numbers for these values will depend on the degree 

4-28
of conservatism desired by the risk assessor and, therefore, should be clearly documented during 

the Risk Characterization phase. 

4-29
5. AQUATIC ECOLOGICAL RISK ASSESSMENT FOR METALS 

This chapter describes how to incorporate the metals risk assessment principles described 

in Chapters 1 and 2 into ecological risk assessments involving aquatic-based receptors. 

Specifically, the following discussion focuses on the relationship between each metal principle 

and components of the EPA’s Framework for Ecological Risk Assessment (U.S. EPA, 1992a) 

and subsequent guidelines (U.S. EPA, 1998a). These components include Problem Formulation, 

Characterization of Exposure, Characterization of Effects, and Risk Characterization. The 

Problem Formulation phase consists of defining assessment endpoints, risk hypotheses, and a 

conceptual model to produce an analysis plan for the risk assessment. Chapter 2 discussed the 

consideration of the metals principles in the Problem Formulation phase. In this chapter 

consideration of the metals principles in the characterization of exposure, effects, and risk to 

aquatic organisms is discussed. 

Consistent with the previously stated scope and purpose of this Framework, not all 

aspects of the ecological risk assessment process are discussed. Only those aspects of the aquatic 

ecological risk assessment process and associated technical issues with the greatest relevancy to 

the metals principles are included. Also emphasized is how the geographic scale (e.g., site 

specific, regional, national) and analytical scope (e.g., screening vs. definitive analysis) of 

aquatic risk assessments affect the extent to which the metals principles can be incorporated. 

Although these principles apply equally to risk assessments involving terrestrial-based ecological 

receptors, many of the methods and tools that can be used to implement these principles differ 

between the aquatic and terrestrial environments. Thus, a separate discussion of how the metals 

principles apply to ecological risk assessment in the terrestrial environment is provided in 

Chapter 6. 

5.1. METALS PRINCIPLES 

Metals have specific environmental and biotic attributes that should be considered in all 

risk assessments. Specifically, these attributes fall into the risk assessment paradigm as follows: 

Background levels Exposure Assessment 

Mixtures Exposure and Effects Assessment 

Essentiality Effects Assessment 

Forms of metals Exposure and Effects Assessment 

Toxicokinetics/toxicodynamics Exposure Assessment (bioavailability) and 

Effects Assessment (absorption, distribution, 

metabolism, and excretion [ADME] and 

toxicity) 

5-1
5.2. CHARACTERIZATION OF EXPOSURE 

The Exposure Characterization phase describes the 

potential or actual contact or co-occurrence of stressors with 

receptors. It includes analysis of stressor sources, their 

distribution in the environment, and the extent and pattern of 

contact or co-occurrence to produce an exposure profile for the 

ecological receptor(s) of concern. Further guidance on 

characterizing exposure in ecological risk assessments is found 

in U.S. EPA (1998a, 1992a). 

Exposure Profile 

The exposure profile should 

describe the exposure 

pathways from stressor source 

to the receptor, the exposure 

intensity, its spatial and 

temporal distribution of cooccurrence, and the impact of 

variability and uncertainty on 

the exposure estimates. 

5.2.1. Background Levels 

Background levels refers to those concentrations of metals that derive from natural as 

well as anthropogenic sources that are not the focus of the risk assessment. In aquatic 

ecosystems, metal concentrations vary widely over space and time owing to differences in 

watershed geology, hydrology, anthropogenic and natural loads from “nontarget” sources, and 

other factors. Depending on the magnitude of the exposure associated with these factors, 

background metal concentrations can account for a significant portion of total metal exposure. 

Furthermore, certain essential metals can bioaccumulate to high levels in some aquatic 

organisms (e.g., Zn in barnacles, Cu in crayfish) due to species-specific physiological 

requirements, regardless of source. Even some nonessential metals can naturally bioaccumulate 

to significant levels as a result of mimicry of essential metals or sequestration and storage. Thus, 

the risk assessor needs knowledge of background concentrations in order to characterize 

exposure and to differentiate risk associated with metal sources already in the environment from 

risk associated with the metal sources of 

concern in the assessment. Environmental Background in Exposure 

Depending on the design and context Assessment 

for the assessment, the risk assessor needs to What are the environmental background 

concentrations at the site(s) of interest? How do 

address several questions and issues they vary over relevant spatial and temporal 

scales? pertaining to background concentrations of What is the relationship between 

metals. Obtaining reliable estimates of environmental background and toxicologically 

relevant metal concentrations? background concentrations can be Can natural and anthropogenic metal be 

challenging, particularly at larger spatial distinguished? 

To what extent are background concentrations 

scales. Risk assessors are cautioned against being extrapolated over space and time? 

using background metals in soils as surrogates What level of confidence (uncertainty) exists in 

the estimate of environmental background 

for sediments due to differences in the concentrations? 

biogeochemical processes between these two 

environments. Metal concentrations in 

5-2
sediments can be impacted by sediment physicochemical composition and localized sediment 

transport processes. Fractionation of sediment cores, in combination with careful estimates of 

sedimentation rates, has been used to distinguish pre- vs. post-industrialized contributions of 

metals via atmospheric transport (e.g., mercury) (U.S. EPA, 1997b). Risk assessors may find 

this type of sediment core analysis useful for differentiating natural levels versus levels 

associated with anthropogenic sources. 

Concentrations of metals in the water column vary over time and are highly responsive to 

hydrological changes. In site-specific risk assessments, the risk assessor may quantify 

background levels by measuring metal concentrations at sites upstream from the area of concern. 

National databases of metal concentrations in various aquatic media (sediments, water, biota) 

include the following: 

• The EPA STOrage and RETrieval (STORET) database (accessible at 

http://www.epa.gov/storet/); 

• The National Sediment Quality Survey (NSQS) (accessible at 

http://www.epa.gov/waterscience/cs/nsidbase.html); 

• Environmental Monitoring and Assessment Program (EMAP) (accessible at 

http://www.epa.gov/emap/); 

• The National Stream Quality Accounting Network (NASQAN) (accessible at 

http://water.usgs.gov/nasqan/), and 

• The Hydraulic Benchmark Network (HBN) (accessible at 

http://ny.cf.er.usgs.gov/hbn/). 

The risk assessor should keep in mind that with the exception of the HBN, these 

databases have not been established to distinguish background concentrations from local, 

anthropogenic sources of metal loadings (e.g., industrial discharges, stormwater runoff) and, 

thus, may reflect significant anthropogenic loadings of metals to environmental media. The 

HBN was established to provide long-term measurements of streamflow and water quality in 

areas that are minimally affected by human activities. While the HBN contains long-term 

measurements of a number of parameters that are known to affect the bioavailability of metals 

(e.g., dissolved organic carbon [DOC], inorganic ions such as Ca and Mg, pH, conductivity), it 

does not contain information on metals of typical regulatory concern, with the exception of Al. 

5.2.2. Forms of Metals 

The physical and chemical forms of metals affect exposure, bioavailablity, and 

subsequent effects and are influenced by physicochemical environmental conditions. Nationallevel assessments involve a broad range of environmental conditions; therefore, the risk assessor 

5-3
should account for different metal species in different locations and water body types. As 

assessments transition from national, to regional, and to local, the assessor should incorporate 

site-specific sediment and water quality parameters that influence metal speciation, 

complexation, and sorption onto biological surfaces (e.g., pH, organic carbon, inorganic ligands, 

Ca, Mg, sulfide). Speciation models (e.g., MINEQL) combined with biotic ligand models offer a 

framework for addressing the differential occurrence and toxicity of various metal forms. Risk 

assessors should be aware of the difficulty in applying this approach to assessments involving 

large regional or national scales because of the variability in model parameter values (including 

covariance among parameters) that occurs across locations. Information about the range of the 

input parameters can be derived from available databases (see Section 5.2.1), and the risk 

assessor can decide what value to use (e.g., minimum, maximum, mean) depending on the 

degree of conservatism desired in the assessment. The risk assessor should include this 

information in the Risk Characterization as part of the overall discussion of assumptions and 

uncertainties in the assessment. Risk assessors can directly assess the metal forms for sitespecific assessments or estimate what these would be based on sediment/water parameters (see 

Section 3.1.3, Environmental Chemistry). 

5.2.3. Exposure Pathway Analysis 

For aquatic organisms, potential routes of exposure to 

metals include absorption across (or in some cases adsorption 

to) respiratory organs, dermal absorption, sediment ingestion, 

and food ingestion. Quantifying exposure and uptake by the 

respitory route is a particular challenge to aquatic risk 

assessors because of the differing types of respiratory organs 

among aquatic species, the dynamic nature of the respiratory 

process in water, and the intimate contact between the 

receptor and metals dissolved in waters. Further 

complicating the issue, some respiratory organs can also be involved in locomotion, excretion, 

ion regulation and the capture, sorting, and ingestion of food. Similarly, risk assessors may find 

the ingestion route difficult to define for aquatic receptors because of the diversity of feeding 

modes and food sources, such as sediments, suspended solids, microflora, animal tissues, and 

plant tissues. The use of stable isotope techniques has contributed greatly to evaluating the role 

of diet in contaminant accumulation (including metals) by precisely defining trophic interactions 

(e.g., Kidd et al., 1995; Jarman et al., 1996). The absorption route can involve uptake across a 

phytoplankton cell membrane, amphibian skin, arthropod exoskeleton, the egg membrane, or the 

integument of an infaunal clam or annelid. 

Despite the complexities associated with quantifying exposure of aquatic animals to 

metals from multiple routes of uptake, risk assessors can find a significant amount of information 

5-4

Exposure Routes 

For aquatic organisms, 

pathways of exposure to 

metals include movement from 

water to sediments (and vice 

versa to a lesser extent) and 

through the food web; air 

deposition directly into aquatic 

systems or through run-off; 

and sedimentation from soils. 
on the relative importance of the different uptake pathways (Wang, 2002; Hook and Fisher, 

2001b; Fisher et al., 1996; Bjerregaard et al., 1985). Applications of one-compartment biokinetic 

models using laboratory-based measurements of key model parameters (assimilation efficiency, 

metal uptake rates from water and food elimination rates) have been extended to field situations 

for populations of a diverse array of aquatic species, including freshwater and marine bivalves, 

various crustaceans such as copepods, amphipods, and crab, aquatic insects, and fish (e.g., 

Luoma and Rainbow, 2005; Stewart and Fisher, 2003; Griscom et al., 2002; Baines et al., 2002; 

Fisher et al., 2000, 1996; Roditi et al., 2000; Wang et al., 1996; Luoma et al., 1992). Sitespecific model predictions for metal concentrations in animal tissues are strikingly close to 

independent field measurements for diverse water bodies, suggesting that it is possible for risk 

assessors to account for the major processes governing contaminant concentrations in aquatic 

animals and that laboratory-derived kinetic parameters are applicable to natural conditions 

(Luoma and Rainbow, 2005). Thus, these models provide tools for risk assessors to use when 

addressing metal exposure and uptake, and they can be used to determine the relative importance 

of different routes of exposure (Landrum et al., 1992; Wang et al., 1996). 

Risk assessors should always consider temporal aspects of exposure, particularly in 

aquatic systems that respond to frequent shifts in hydrology. Rapid speciation and phase 

changes associated with changes in pH/Eh make temporal issues particularly germane to metals. 

Fluctuating or pulsed exposures occur in situations such as rapid changes in pH/Eh associated 

with photosynthesis and respiration, hypolimnetic discharge from stratified reservoirs, biocide 

(e.g., copper sulfate) spraying, ingestion of prey items with seasonally high metal concentrations, 

surface waters receiving wastewater treatment plant effluent, urban storm water, snowmelt, and 

acid precipitation runoff. Transient metal concentrations may be orders of magnitude higher 

than typical or average concentrations but last for only a few hours. These episodic exposure 

scenarios have been poorly characterized for metals (Butcher et al., 2006). Any risk assessment 

for metals should clearly state all assumptions about duration of exposure and what uncertainties 

are added to the risk model as a consequence. 

5.2.4. Fate and Transport of Metals 

Risk assessors routinely use transport and fate computational models to describe and 

quantify exposure pathways. Models also are useful in situations where risk assessors need an 

estimate of future exposure levels that are expected to result from the implementation of some 

permitting action or remediation measures at local, regional, or national scales. Numerous 

models are available for use; most are based on the same fundamental principles. Metals are 

ubiquitous in the environment and within each media compartment they are present in 

association with air, water (freely dissolved metal or as organic and inorganic metal complexes), 

and particles (sorbed, precipitated, or incorporated within a mineral phase). The risk assessor 

can find a detailed discussion of the fate and transport of metals in Section 3.2. No single, 

5-5
currently available model includes all the desirable metal-specific features for aquatic systems. 

Discussions of the family of available aquatic transport and fate models, as well as a number of 

chemical equilibrium models, may be found in Paquin et al. (2003). 

5.2.5. Toxicokinetics and Toxicodynamics (Bioavailability and Bioaccumulation Issues) 

5.2.5.1. Aqueous Phase 

In the dissolved phase, metals can exist as free ions as well as in a variety of complexed 

forms. These forms, or species, are of key importance in understanding bioavailability, and the 

hazard and risk assessments of waterborne metals are complicated by the fact that metal species 

differ in their toxicological properties. For many metals in aquatic systems, it is the free ionic 

form that is most responsible for toxicity. For example, Cu2+ has been directly linked to toxicity 

in fish and invertebrates while Cu complexed by dissolved organic matter does not induce 

toxicity to the same degree (Ma et al., 1999; Erickson et al., 1996) because of its reduced 

availability for uptake by the organism. However, the risk assessor should be aware that 

although toxicity of metals bound to DOC is reduced, it is not eliminated entirely and can 

contribute to the total metal loading to an ecosystem and subsequent toxic effects (McGeer et al., 

2002; Erickson et al., 1996). On the other hand, there are cases where nontoxic metal species are 

bioavailable and taken up by the organism but cause no adverse response (e.g., Ag-Cl complexes 

in rainbow trout) (McGeer and Wood, 1998). Risk assessors should recognize that the presence 

of metal within an organism cannot always be used as a surrogate for toxic response. 

Risk assessors can choose among a variety of methods to account for relative 

bioavailability of metals in aquatic systems, including hardness adjustments, water-effect ratio 

(WER), Free Ion Activity Model (FIAM), and aquatic Biotic Ligand Models (BLMs) (Paquin et 

al., 2002a). Each method contains strengths and limitations and may not be amenable to all 

types of assessments (e.g., ranking/classification, national, and site-specific assessments). For 

example, adjustment of aqueous metal concentrations for differences in water hardness was 

among the first computational methods to account for bioavailability differences between the 

laboratory and the field when applying EPA water quality criteria. Although these adjustments 

are relatively easy to apply, they require empirical data to define the toxicity-water hardness 

relationship. Thus, they are more amenable to site-specific risk assessments, although even in 

those cases they do not account for other water quality factors that affect bioavailability (e.g., 

DOC, pH). The water hardness approach has been applied at the national level through a 

statement of water quality criteria as hardness-based equations rather than as single values. Risk 

assessors can choose to use ranges, means, or median values when conducting large-scale (e.g., 

regional) assessments, but they will need to acknowledge this uncertainty during the risk 

characterization phase. 

5-6
Bioavailability adjustments using WERs incorporate the combined effects of all water 

quality parameters present in site water on bioavailability of metals of interest relative to what 

was measured in laboratory tests (U.S. EPA, 1994c). Thus, compared to hardness adjustments, 

WERs encompass a broader array of water quality factors that can impact bioavailability. 

However, WERs are relatively resource intensive (requiring toxicity testing), are applicable only 

on a site-specific basis, and are not easily adjusted to account for temporal or spatial variability. 

This relationship between speciation and bioavailability also has been explained through 

the free ion activity model (FIAM, Campbell, 1995). This model produces speciation profiles of 

a metal in an aquatic system and provides insight into the relative bioavailabilities of the 

different forms of metal as well as the importance of complexation. Models available for the 

calculation of metal speciation in natural waters are reviewed in Section 3.1.5, Sediment 

Chemistry, and include MINEQL (Schecher and McAvoy, 1994; Westall et al., 1976), 

MINTEQA2 (Brown and Allison, 1987), CHESS (Santore and Driscoll, 1995), WHAM 

(Tipping, 1994), and PHREEQ (Parkhurst et al., 1980). The risk assessor should review Paquin 

et al. (2003) for a more in-depth understanding of these models and how to select among them 

for particular places and types of assessments. 

The BLM approach successfully combines the influences of speciation (e.g., free metal 

ion, DOC complexation) and cationic competition (e.g., K+, Na+, Ca+2, Mg+2) on metal toxicity 

in fish (De Schamphelaere and Janssen, 2004, 2002; De Schamphelaere et al., 2004, 2003, 2002; 

Heijerick et al., 2002a, b; Di Toro et al., 2001; Santore et al., 2001; McGeer et al., 2000). The 

model can be used to distinguish, at least conceptually, metals that will bioconcentrate at the site 

of toxicity (e.g., gill or other biotic ligand) from the total metal pools in an organism and the 

bioavailable metal pool in the exposure media. The model also can be applied to algae (De 

Schamphelaere et al., 2003; Heijerick et al., 2002) and Daphnia (De Schamphelaere and Janssen, 

2004, 2002; De Schamphelaere et al., 2004, 2002). The BLM recently has been incorporated 

into draft revisions to EPA’s national water quality criteria for Cu (it has been used in risk 

assessments at a range of geographic scales), and it is being applied as an alternative to the WER 

approaches for setting site-specific discharge objectives. The BLM has the potential to address 

spatial and temporal factors that affect bioavailability, provided that the variability in water 

quality parameters used as inputs to the model can be quantified or predicted (e.g., pH, DOC, K+, 

Na+, Ca+2, Mg+2). However, the risk assessor should be aware of the many limitations in 

applying the BLM (or any Free Ion model). For example, the development of the BLM has 

focused primarily on bioavailability and acute exposures. Work has begun to extend the BLM to 

chronic toxicity for some organism/metal combinations (e.g., De Schamphelaere and Janssen, 

2004; Paquin et al., 2002a, b), and further development is expected. Also, the BLM is currently 

based on metal uptake through the dissolved phase; thus, additional research is needed to address 

metal uptake and toxicity via the diet. 

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5.2.5.2. Sediment Phase 

Risk assessors have several approaches for 

estimating exposures to sediment-associated metals 

that account for bioavailability differences. The 

equilibrium partitioning approach (EqP) assumes that

chemical activity in the sediment, as indexed by 

chemical concentration in the interstitial water, is 

proportional to the chemical’s bioavailability to 

sediment-dwelling organisms. In anoxic sediments, sulfides provide the primary binding phase 

for many cationic metals. These metal sulfides are highly insoluble and are thought to have very 

low toxicity. Thus, in sediments where there is more sulfide than metal, most cationic metals 

should be present as insoluble sulfides and relatively nontoxic. The amount of reactive sulfide is 

quantified by measuring the amount of sulfide freed when sediment is extracted with 1 N HCl. 

This procedurally defined quantity is known as acid volatile sulfide (AVS). The amount of 

reactive metal is determined from the same extraction by measuring the metal concentration in 

the acid extract. This quantity is known as simultaneously extracted metal (SEM). The risk 

assessor then can determine the potential bioavailability of a metal by comparing the relative 

molar concentrations of the SEM and AVS. When SEM-AVS < 0, sufficient sulfide exists to 

bind all SEM, and metal toxicity is not expected. When SEM-AVS > 0, metal is present beyond 

the binding capacity of sulfide, and toxicity may occur if there is sufficient excess metal but not 

sufficient other binding phases to bind the metal. Use of this SEM-AVS as exposure estimates 

that are correlated with toxicity of metals in sediment has been explored closely for many metals 

(Ankley et al., 1996, 1991; Berry et al., 1996; Hansen et al., 1996; Carlson et al., 1991; Di Toro 

et al., 1990). 

However, risk assessors should be aware that although the correspondence of SEM-AVS 

to toxicity was found, some questions remain about the applicability of the approach to all 

benthic organisms because it is based on the chemistry of bulk anoxic sediment, and many 

organisms live in oxygenated burrows. In addition, several studies have shown some degree of 

metal accumulation in organisms exposed to sediments where sulfide is in excess and metals are 

thought to be nonbioavailable, or at least nontoxic (Ankley et al., 1996). However, the lack of 

toxicity observed when AVS exceeds SEM suggests that this bioaccumulated metal may not be 

biologically available. A better understanding of the mechanisms of metal accumulation from 

sediment and their relationship to toxic effects is needed to help interpret these issues. Until 

such information becomes available, risk assessors can use the SEM-AVS model in exposure 

estimations as long as its shortcomings are acknowledged appropriately and uncertainties are 

recorded in the Risk Characterization phase of the assessment. 

Other tools risk assessors can use to determine the bioavailable concentrations of 

Sediment-Associated Metals 

When the molar concentrations of 

acid-volatile sulfide in sediment 

exceed the amount of simultaneously 

extracted metal, the metals are 

expected to be associated with the 

solid phase and not to be toxic. 

sediment-bound metals include metal concentrations in the chemical extracts (Fan and Wang, 

5-8
2001; Babukutty and Chacko, 1995; Tessier et al., 1984), acid extracts (Langston, 1980; Luoma 

and Bryan, 1978), or biomimetic extracts (Weston and Maruya, 2002; Mayer et al., 2001; Chen 

and Mayer, 1998). However, no consensus yet exists on their best use for different types of 

metals or metalloids. Several other methods have been proposed. Based on the premise that iron 

oxides in oxic sediments lower metal bioavailability, Fe in a 1 N HCl sediment extract has been 

used to normalize metal exposure concentrations (Luoma and Bryan, 1978). Increasing 

concentrations of organic carbon can decrease metal bioavailability (Crecelus et al., 1982), so 

normalization of sediment metal concentrations to organic carbon content has been conducted in 

other cases. The more readily extracted metals from sequential chemical extraction schemes 

tend to be the most bioavailable (Young and Harvey, 1991; Tessier et al., 1984) and have been 

used to estimate bioavailable metal. 

5.2.5.3. Dietary Phase 

As discussed in Section 5.2.3 (Exposure Pathway 

Analysis) and illustrated by the conceptual model for 

bioavailability (Figure 2-2), it is well established that 

dietary exposure to metals can result in accumulation of 

metals in aquatic organisms. What is less well 

established is the best way to express dietary exposure in 

a way that can be linked to potentially toxic effects (eithe

directly to the aquatic organism or its predator). The 

main reason for this ambiguity is that the bioavailability 

of dietary metals varies widely across organisms and exposure conditions and standardized 

approaches are not available for predicting toxicity. The subsequent discussion elaborates on 

this point and provides some suggestions for how risk assessors might address dietary metals in 

different assessment contexts (screening vs. definitive). 

After ingestion, some of the dietary metal can be released from the ingested particle into 

the gastrointestinal fluids of the animal (Chen and Mayer, 1999; Mayer et al., 1997; Gagnon and 

Fisher, 1997) and become available for assimilation into the tissues of the animal and the tissues 

of its consumer (i.e., trophic transfer). Assimilation efficiency (i.e., the net amount of metal 

retained in tissues relative to the amount ingested from food) is a common measure of the 

bioavailability of a chemical from food, and the risk assessor may find this to be a useful 

parameter for comparing the potential for toxicity among different types of organisms. 

Assimilation efficiency is also an important input parameter for estimating metal 

bioaccumulation using kinetic-based bioaccumulation models (e.g., Luoma and Rainbow, 2005). 

Assimilation efficiencies can vary widely depending on the metal, its form and distribution in 

prey, species digestive physiology (e.g., gut residence time), environmental conditions (e.g., 

temperature), food quality, food ingestion rate, and metal concentration in the diet. Thus, risk 

Dietborne Metal Exposure 

Risk assessors should consider 

dietborne metal exposure in two 

contexts: (1) dietborne exposure 

leading to accumulation and 

exposure to higher levels in the food 

chain (e.g., humans, wildlife) and (2) 

dietborne exposure leading to direct 

r effects on exposed organisms. 

5-9
assessors should consider likely ranges of assimilation efficiencies for a particular metal-animal 

combination when evaluating metal bioavailability from the diet. A number of reviews have 

summarized current knowledge of assimilation efficiencies of ingested metals among different 

aquatic animal species (Wang and Fisher, 1999; Fisher and Reinfelder, 1995). 

The distribution and form of metals in dietary organisms is of critical importance for 

understanding the bioavailability of dietary metals and trophic transfer potential. For example, 

metals in the cytosolic fraction of phytoplankton and “soft parts” of zooplankton have been 

shown to correlate well with bioaccumulated metal in their predators (e.g., herbivores and 

plantivorous fish, respectively) (Wang and Fisher, 1996; Reinfelder and Fisher, 1994a, b). 

Metals sorbed to the cell wall of phytoplankton and the exoskeleton of zooplankton were poorly 

assimilated by consumers. The bioavailability of metal-enriched granules in prey (a 

detoxification and storage mechanism exhibited by some organisms) has been shown to be 

negligible or substantially reduced when consumed by certain predators (e.g., Wallace et al., 

2003, 1998; Wang, 2002; Wang and Fisher, 1999; Nott and Nicolaidou, 1990). However, the 

risk assessor should be aware that the bioavailability of metal-enriched granules in prey can vary 

among metals and with type of granule (Wang, 2002; Mason and Jenkins, 1995; Nott and 

Nicolaidou, 1990) and may also depend somewhat on digestive physiology of the predator (e.g., 

gut pH, retention time). As a result of these findings, fractionating body burdens of metals (e.g., 

cytosolic metal vs. metal granules) has been suggested as a better means of identifying the 

bioavailable fraction of dietborne metals (Seebaugh and Wallace, 2004; Wallace et al., 2003; 

Fisher and Reinfelder, 1995; Reinfelder and Fisher, 1994). Although such techniques show 

promise for operationally defining the extent to which dietary metals may be bioavailable for 

trophic transfer in aquatic food webs, risk assessors should understand that broad-scale 

application of these techniques to metals risk assessments is presently limited by the relatively 

small number of metals and predator-prey relationships evaluated. 

Despite the uncertainties associated with bioavailability and trophic transfer of dietary 

metals, the use of whole-body inorganic metal concentrations in prey species may have some 

utility to risk assessors for conservatively screening for exposure and potential risks to 

consumers (i.e., in cases where whole-body residues 

are below dietary toxic thresholds). For more 

definitive assessments, further research is needed to Trophic Transfer 

quantify the bioavailability and effects of inorganic Trophic transfer is the transfer of a 

dietary metals, with the exception of certain chemical from prey species to a predator 

species via dietary exposure. 

organometallics (e.g., methyl mercury) and Biomagnification is a type of trophic 

transfer where chemical concentrations 

metalloids (e.g., Se) where dietary toxicity has been increase in organisms from a lower 

well established. trophic level to a higher trophic level 

within the same food web. 

Biodilution represents a decrease in 

organism concentration with increasing 

trophic level. 

5-10
5.2.5.4. Bioaccumulation and Trophic Transfer 

Assessing and predicting the bioaccumulation of metals in aquatic ecosystems is a 

component of many Agency regulatory and nonregulatory activities (e.g., chemical 

ranking/classification, derivation of national water quality criteria, Superfund site risk 

assessments; see Section 2.3.1 for definitions and a conceptual model related to 

bioaccumulation). Interest in metals bioaccumulation originates from concerns regarding the 

direct impact of metals on organisms accumulating the metal and indirect impacts on their 

consumers (i.e., trophic transfer). Unlike certain persistent and bioaccumulative organic 

compounds, which tend to biomagnify in aquatic food webs (e.g., DDT/DDE, PCBs, 

PCDD/PCDF), inorganic metal compounds rarely biomagnify across three or more trophic levels 

(McGeer et al., 2004; Suedel et al., 1994); however, certain organometallics can biomagnify in 

aquatic food chains. Some metals (e.g., lead) tend to biodilute in aquatic food webs. Risk 

assessors should not interpret lack of biomagnification as lack of exposure or concern via trophic 

transfer (see text box). Even in the absence of biomagnification, aquatic organisms can 

bioaccumulate relatively large amounts of metals and become a significant source of dietary 

metal to their predators (Reinfelder et al., 1998). 

For many nonionic organic chemicals, risk assessors can derive first-order 

approximations of bioaccumulation potential from information on chemical properties and 

organism attributes (e.g., Kow and lipid content) and their use as inputs to simplified, fugacitybased models (e.g., Gobas et al., 1993). For metals, analogous methods to predict 

bioaccumulation based on simple chemical properties are not available or are not widely 

validated. The lack of analogous models for metals is likely due, at least in part, to the high 

degree of specificity exhibited by the mechanisms and processes underlying metals 

bioaccumulation (e.g., speciation, exposure conditions, and organism physiology) (see McGeer 

et al., 2004; Rainbow, 2002; Mason and Jenkins, 1995). As a result, risk assessors currently are 

limited to using an empirical approach for assessing and predicting metals bioaccumulation. 

Typically, this requires direct measurement of metal concentrations in the organism or 

experimentally-determined parameters for use as input to bioaccumulation models (e.g., gill 

uptake rate, elimination rate, assimilation efficiency). 

Aquatic ecological risk assessors commonly use bioconcentration factors (BCFs) and 

bioaccumulation factors (BAFs) to quantify chemical accumulation in tissue relative to 

concentration in water. BCFs and BAFs are determined as the ratio of the chemical 

concentration in tissue to its concentration in water (using the steady-state method) or as a ratio 

of uptake rate (ku) and elimination rate (ke) constants (using the kinetic method). Measurement 

of BCFs or BAFs usually is conducted for conditions that approximate steady-state (i.e., where 

accumulation remains relatively constant due to chemical uptake being offset by its elimination 

by the organism). It is assumed that the greater the BAF or BCF, the greater a chemical’s 

bioaccumulation or bioconcentration potential. 

5-11
Risk assessors should recognize that considerable uncertainty can be associated with the 

application of literature-derived BCFs and BAFs for assessing the risks of metals, as variability 

in BCFs and BAFs for metals is known to be high (e.g., 50-fold or higher within a metal). Most 

of this uncertainty results from bioavailability differences between the studies from which the 

BCF or BAF is measured and the site(s) to which it is being applied (e.g., water quality 

characteristics, metal speciation, exposure pathways). 

Other sources of uncertainty that risk assessors should consider in the broad application 

of BCF/BAF data are rooted in the complex mechanisms of metal toxicokinetics (uptake, 

metabolism, distribution, elimination). For example, unlike hydrophobic, nonionic organic 

chemicals where uptake across biological membranes generally occurs via passive diffusion, the 

uptake of metals is believed to involve a number of specific transport mechanisms. Some of 

these transport mechanisms involve binding with membrane carrier proteins, transport through 

hydrophilic membrane channels, and endocytosis. Passive diffusion is thought to be reserved for 

certain lipid soluble forms of metals, such as alky-metal compounds and neutral, inorganically 

complexed metal species (e.g., HgCl2

0

). The implication of these specific transport mechanisms 

is that metal bioaccumulation can involve saturable uptake kinetics, such that BCFs and BAFs 

depend on exposure concentration. The existence of saturable uptake mechanisms, the presence 

of significant amounts of stored metal in organisms, and the ability of some organisms to 

regulate bioaccumulated metal within certain ranges are all thought to be responsible for the 

inverse relationship that has been frequently reported between BCFs/BAFs and metal exposure 

concentrations (McGeer et al., 2003; Borgman et al., 2004). In these cases, higher BCFs or 

BAFs are associated with lower exposure concentrations and also can be associated with lower 

tissue concentrations within a given BCF or BAF study. This is counter to the implicit 

assumption that higher BCFs or BAFs indicate higher metal hazard. 

As a result of the aforementioned uncertainties, risk assessors must be careful in broadscale application of BCF and BAF data for metals. Specifically, the current science does not 

support the use of a single, generic threshold BCF or BAF value as an indicator of metal hazard. 

Similarly for national risk assessments, use of a single BCF or BAF value holds little utility due 

to high uncertainty that results from differences in bioavailability, exposure conditions, and 

species-specific factors that influence metal bioaccumulation by aquatic organisms. When 

extrapolation across sites is necessary and limited data prevent application of alternative 

approaches (discussed below), uncertainty in the use of BCFs and BAFs can be reduced by 

expressing them as a function of media chemistry (i.e., to address bioavailable metal), exposure 

concentrations (i.e., to address concentration dependency issues), and limiting extrapolations to 

within a particular species or closely related species. The use of BCFs and BAFs for metals 

assessments appears to have most value for site-specific applications, when appropriate 

measurements are taken from the site(s) of interest and extrapolation of BCF/BAF values across 

differing exposure conditions and species is minimized. 

5-12
Risk assessors should be aware of several alternatives for assessing metals 

bioaccumulation that address some of the concerns listed above. One of these is to develop 

regression relationships between tissue and exposure concentrations. Such regression 

relationships have been used to characterize bioaccumulation of metals by soil organisms (U.S. 

EPA, 2003c; Sample et al., 1999), but they have not yet been compiled for aquatic systems. The 

advantage of this technique is that it addresses the dependency of BCF or BAF on exposure 

concentration. However, it does not explicitly adjust for bioavailability differences that occur 

across sites. Another alternative is to use a kinetic-based model for describing bioaccumulation 

(Luoma and Rainbow, 2005; Wang and Zauke, 2004; Kahle and Zauke, 2003; Chang and 

Reinfelder, 2002; Reinfelder et al., 1998). These models can improve predictions of metal 

bioaccumulation in aquatic organisms because they incorporate different exposure routes (e.g., 

water vs. diet) and the dynamic nature of metal bioaccumulation processes. For example, Luoma 

and Rainbow (2005) reviewed the DYNBAM model (a single-compartment, kinetic-based 

bioaccumulation model) and found it to accurately predict metal bioaccumulation for a wide 

range of metals, organisms, and habitats based on data derived from 15 separate studies. 

Importantly, DYNBAM and similar such models require experimental data measured under 

environmentally-realistic conditions in order to derive model parameters for each metal-species 

combination (e.g., uptake and elimination rates, assimilation efficiency, food ingestion rates). 

Compilations of such data on model input parameters are available for some species and metals 

(e.g., Wang and Fisher, 1999, for aquatic invertebrates). Clason et al. (2004) and Kahle and 

Zauke (2003) have developed two-compartment bioaccumulation models for amphipod 

crustaceans that incorporate background metal and saturation of uptake kinetics. Currently, 

however, these models include only the dissolved phase and do not account for uptake from the 

diet. 

The bioaccumulation models described above offer strong promise for improving 

bioaccumulation predictions in aquatic risk assessments for metals and should be considered by 

risk assessors. However, risk assessors should be aware of their limitations. For example, they 

currently do not account for differential partitioning and bioavailability of metal in organisms 

(see Figure 2-2). Empirical methods are being developed to predict metal compartmentalization 

in tissues of aquatic organisms (e.g., Wallace and Luoma, 2003; Wallace et al., 2003), but these 

have not been incorporated into bioaccumulation models. These models also do not explicitly 

address the impact of metal speciation on bioaccumulation or link bioaccumulated metal to toxic 

effects, although such models are under development (Paquin et al., 2002b). Regardless of the 

type of bioaccumulation model used, reductions in uncertainty in metals bioaccumulation 

assessments should be directed at achieving robust connections between the 

bioaccessible/bioavailable form(s) of metals in various exposure media, their accumulation, 

metabolism, and distribution in tissues, and the form(s) of metals that exert their toxicity directly 

to the organism or indirectly to its consumers. 

5-13
5.3. CHARACTERIZATION OF EFFECTS 

5.3.1. Essentiality 

Essentiality refers to the nutritional requirements of an organism for normal metabolic 

function. A key difference between metals and organic chemical contaminants is that some 

metals are required either as macronutrients (e.g., Fe, Ca, Mg) or micronutrients (e.g., Cu, Zn, 

Ni) to maintain a healthy organism. Table 6-3 in Section 6.2.1 classifies the metals addressed in 

this Framework by their known essentiality to plants and animals. 

Consideration of essentiality by the risk assessor is important for several reasons. First, 

the risk assessor should ensure that toxic effects thresholds calculated from an assessment are not 

lower than the nutritional requirements for the particular plant or animal species being evaluated. 

As discussed in Sections 4 and 6, the risk assessor should be aware that such elements exhibit 

classic bell-shaped or biphasic dose-response (or exposure effect) relationships, with adverse 

effects occurring at both high and low concentrations and an optimal mid-range dose or exposure 

(see Abernathy et al., 1993; Chapman and Wang, 1998). For aquatic organisms, information 

about nutritional requirements is available for many commonly tested or cultured species. 

Essentiality issues also impact the bioaccumulation and toxicity of metals since organisms have 

evolved various mechanisms to maintain homeostasis of essential metals. Such mechanisms 

may also impact the bioaccumulation and toxicity of nonessential metals, particularly those that 

share similar binding and uptake mechanisms. In these cases, accumulation is nonlinear with 

respect to exposure concentration, whereby greater uptake and retention of metals occurs at low 

concentrations and uptake rates decrease as exposure media concentrations increase. The impact 

of homeostatic and other mechanisms on bioaccumulation has been discussed in Section 5.2. 

5.3.2. Toxicokinetics/Toxicodynamics (Toxicity Issues) 

For organometallic compounds such as organo-selenium and methyl mercury, toxicity 

from dietary exposure has been shown to contribute substantially to ecological risk at 

environmentally-realistic concentrations and thus should be considered by the risk assessor when 

characterizing the effects of these compounds. Beyond those two organo-metal compounds, 

however, the importance of exposure to dietary metals is much less clear. Toxicity to aquatic 

organisms from dietary exposure to metals has been demonstrated where exposure is sufficiently 

high, although, in some cases, these concentrations are extreme (e.g., 10,000 g/g Cu) (Handy, 

1993). In such cases, it is not clear that this pathway will drive ecological risk, as the 

environmental concentrations necessary to produce these exposures may be so extreme that 

ecological risk will occur first via other pathways (e.g., direct toxicity of waterborne metal). 

In other studies, however, effects from dietary exposure have been demonstrated at 

relatively low exposure concentrations (e.g., zooplankton studies by Hook and Fisher, 2002, 

5-14
2001a, b). This raises additional concern for metals assessment because it increases the potential 

for toxicologically significant exposures to occur in cases where risk via a waterborne pathway is 

low. However, other studies with the same organisms and metals, but somewhat different test 

methods, reached different conclusions regarding the significance of dietary exposure (e.g., De 

Schamphelaere and Janssen, 2004; Meyer et al. [in press]). Dietary exposure of aquatic 

organisms to metals is an active area of research, and it is likely that new data and insights will 

result in a more comprehensive understanding of dietary effects. Until that time, risk assessors 

should make decisions regarding potential risks of dietary metal exposure on a case-by-case 

basis. 

5.3.3. Metal Mixtures 

Mixtures of metals (including metalloids and organic substances) are commonly 

encountered in the natural environment as a result of anthropogenic inputs and should be 

considered by the risk assessor for all assessments. Metal interactions, according to Calamari 

and Alabaster (1980), occur at three levels: 

• Chemical interactions with other constituents in the media 

• Interactions with the physiological processes of the organism, and 

• Interactions at the site of toxic action. 

The joint action of metal mixtures may be expressed in different ways, including 

increasing or decreasing the toxicity relative to that predicted for individual components. As a 

result, the toxicity of metal mixtures has important consequences for metals risk assessments. 

For example, toxicity has been observed for mixtures of metals present individually at nontoxic 

levels (e.g., at levels corresponding to water quality guidelines) (Enserink et al., 1991; Spehar 

and Fiandt, 1986). Despite the importance of considering the effects and mixtures of metals to 

aquatic organisms, risk assessors will find that predicting the toxicity of metal mixtures has 

proven to be a difficult challenge in aquatic toxicology. 

Much of the difficulty in predicting the toxicity of metal mixtures to aquatic organisms 

results from differences in the bioavailability and/or methods used to define the bioavailable 

fraction among toxicological studies and subsequent ambiguity in interpreting mixture toxicity 

data. As discussed in Chapter 2, the bioavailability of metals depends on a suite of factors that 

can affect their speciation, complexation with ligands, and interaction with biological systems 

(e.g., pH, DOC, inorganic anions, and cations). Apart from bioavailability differences, the joint 

action of metal mixtures to aquatic organisms (i.e., antagonism, additivity, synergism) has been 

reported to depend on other aspects of toxicity test design, including the degree of toxicity 

associated with the overall mixture concentration (Mowat and Bundy, 2002; Fargašová, 2001; 

Herkovits et al., 1999; Spehar and Fiandt, 1986), the relative proportion of constituent 

5-15
concentrations (Norwood et al., 2003; Sharma et al., 1999), the duration of the exposure (Marr et 

al., 1998), and several other factors related to experimental design (Norwood et al., 2003). In a 

similar review, the European Centre for Ecotoxicology and Toxicology of Chemicals concluded 

that the acute and chronic toxicity to aquatic organisms of mixtures of metals could not be 

reliably predicted or generalized although they recommended that, in the interim, assuming 

additive effects is likely “a balanced approach” for acute toxicity of metal mixtures (ECTOC, 

2001). 

Given these difficulties in evaluating mixtures effects, risk assessors commonly use two 

simplifying models: concentration addition and effects (response) addition. These models are 

used to classify the combined effects of chemical mixtures as being antagonistic, additive, and 

synergistic (also referred to as “less than additive,” “strictly additive,” and “more than additive,” 

respectively). Both models use metal concentrations in media to generate concentrationresponse curves for individual metals, and these data are then used to generate specific critical 

concentrations for mixture models. In the concentration addition model, all metals in a mixture 

are added together to predict toxicity; differing potencies are taken into account by converting 

chemical concentrations to an equitoxic dose, such as toxic units (TUs) or toxicity equivalence 

factors (TEFs), which converts all metals to one metal concentration. Concentration addition is 

often used when the constituents are known or assumed to act through the same or similar MOA. 

However, risk assessors should be aware that applying the concentration addition model to 

mixtures containing many metal constituents (particularly those well below toxic levels) can 

result in an upward bias in predicted mixture toxicity (Newman et al., 2004). 

In the effects addition model, differing potencies are ignored, and the effect of each 

metal’s concentration in a mixture is combined to predict mixture toxicity. The effects addition 

model is used when constituents act or are assumed to act independently (i.e., different MOAs). 

Thus, the risk assessor defines the nature of the metals’ joint action (i.e., independent or similar 

MOA) to decide when to apply either the concentration addition or effects addition model. The 

risk assessor should consult information on the MOA, capacities to act as analogues for other 

metals, essentialities, and ligand binding tendencies to choose among these types of models. 

The assumption of additivity has some regulatory precedence for use when addressing the 

toxicity of mixtures although not necessarily for metals. For example, the concentration addition 

approach is recommended for use by Australia and New Zealand for evaluating whether a 

mixture of less than six constituents exceeds their water quality guidelines (ANZECC and 

ARMGANZ, 2000). Similarly, additivity is assumed by EPA when evaluating the combined 

acute toxicity of multiple toxic effluents on the basis of whole-effluent toxicity data (U.S. EPA, 

1991). Additivity is not assumed for chronic exposures due to lack of supporting data. For 

predicting the direct toxicity of mixtures of cationic metals in sediments to benthic organisms, 

EPA uses the ∑SEM-AVS approach described previously (see Section 5.2.5 Toxicokinetics and 

Toxicodynamics (Bioavailability and Bioaccumulation). Note, however, that this method is 

5-16
limited to six cationic metals (Cd, Cu, Pb, Ni, Ag, and Zn). Furthermore, this method is 

considered a “no effect guideline,” whereby the absence of toxicity can be predicted reliably 

(when ∑SEM<AVS) but the occurrence of toxicity (when ∑SEM>AVS) cannot be because of 

other factors that are not accounted for, which reduce metal toxicity. 

Risk assessors should also consider the QICAR approach (described in Section 3.1.1) for 

addressing the toxicity of metal mixtures. Unsatisfied with the qualitative conclusions of 

Newman and McCloskey (1996), Ownby and Newman (2003) fit binary metal mixture data 

derived from the Microtox assay to develop a model of joint independent action (Finney, 1947). 

They predicted that the joint action of combined metals will increasingly deviate from 

independent action as their ligand-binding chemistries become more and more similar. Although 

Microtox is considered to be a useful tool for organic contaminants, its sensitivity for evaluating 

metal toxicity has been called into question (Willemson et al., 1995). 

It is possible that receptor binding models (e.g., FIAM) may be expanded in the future to 

include mixtures. In theory, if two metals compete for binding to the same site of toxic action, it 

should be possible to model the total metal bound to that site and, hence, to predict metal toxicity 

using a mechanistic receptor binding approach in an effects addition model. Alternatively, if two 

metals do not compete for the same binding site, then these models may provide more reliable 

estimates of individual metal bioavailability, which then can be combined in more accurate 

effects addition models. However, at present, these possibilities remain theoretical. 

Furthermore, this approach, while improving the ability to assess the effects of metal mixtures, 

does not include temporal aspects (i.e., “time-to-response” versus concentration). 

From the preceding discussion, it should be clear that the accurate prediction of the joint 

toxicity of metal mixtures to aquatic organisms remains a significant challenge for the risk 

assessor. For site-specific assessments, risk assessors are encouraged to assess mixture toxicity 

using in situ measurements (i.e., bioassays using site water or sediments). This approach is the 

foundation of the WER procedures used by EPA for making site-specific bioavailability 

adjustments to metals criteria. For site-specific assessments involving sediments, risk assessors 

should consider using the ∑SEM-AVS approach as a no-effect threshold. For national-level 

assessments, there is some precedence for assuming additive toxicity of mixture constituents, 

particularly when considering acute effects. However, the risk assessor should carefully consider 

the limitations to assuming strict additivity (i.e., potential for overprediction or underprediction 

of toxicity) and highlight these uncertainties in the Risk Characterization phase of the aquatic 

risk assessment. 

5.3.4. Critical Body Residues 

The bioavailability of metals from multiple exposure routes (water column, food, 

sediments) should be considered in aquatic risk assessments to account for relative contributions 

to overall toxicity. In concept, expressing toxicity on the basis of tissue residues is an attractive 

5-17
approach to accomplishing this because it integrates chemical uptake from different routes of 

exposure, accounts for differences in bioavailability from exposure media, and addresses 

differences in toxicokinetics that occur for different species. 

Expressing toxicological effects on the basis of internal (tissue) concentrations (e.g., use 

of critical body residues [CBR] or residue-response relationships) has gained significant 

attention in the aquatic ecotoxicology literature, particularly for organic chemicals (e.g., 

Landrum et al., 2005, 2004; Escher and Hermens, 2002; McCarty and Mackay, 1993; Cook et 

al., 1993, 1989; McCarty, 1986; Veith et al., 1983; Könemann, 1981). For many nonionic 

organic chemicals, available data indicate that whole-body burdens of chemical (normalized to 

lipid content) can serve as useful metrics of toxicological dose, and these relationships appear to 

be independent of whether exposure was via water or diet. A major strength of the CBR 

approach for organic chemicals is that it effectively integrates different exposure pathways into a 

single expression of dose and toxicological potency. 

For metals (aside from organo-selenium and methyl mercury), the situation is far more 

complex and the CBR approach does not appear to be a robust indicator of toxic dose. One 

reason why the CBR approach currently appears more limited for metals relates to differences in 

their mechanisms of uptake, distribution, and disposition in aquatic organisms. Specifically, the 

distribution of nonionic organic chemicals in organisms is largely influenced by passive 

partitioning. In contrast, the uptake, distribution, and disposition of metals are typically 

governed by highly-specific biochemical processes that alter the metal form and involve 

facilitated or active transport. For example, some organisms take up metal and sequester it into 

“storage” compartments in chemical forms that have little toxicological potency, whereas other 

organisms actively excrete excess metals. As a basis for improving residue-response 

relationships, some studies have suggested that the metal concentration in the cellular cytosol (as 

opposed to that bound to cell walls or sequestered in nonbioavailable metal granules) may 

provide a better expression of internal metal dose associated with toxic effects (Wallace and 

Luoma, 2003; Wallace et al., 2003; Wallace and Lopez, 1996). 

Other researchers have suggested that CBR relationships are confounded because the 

factor that determines the effects is not whole-body concentration per se, but the rate of metal 

uptake in relation to metabolic capacity for detoxification and storage; therefore, the effects are 

governed by factors that influence the rate of uptake. When uptake rate is elevated, the 

concentration of metabolically active metal at the site(s) of action increases (e.g., the spillover 

hypothesis) and effects ensue (Rainbow, 2002). Because different species of aquatic organisms 

invoke different “accumulation strategies” (i.e., involving combinations of regulation, 

detoxification, and storage), considerable difficulties arise among species when interpreting the 

toxicological significance of metal whole-body residues. 

Therefore, risk assessors should ensure that a toxicologically valid residue-response 

relationship supports the CBR threshold before using tissue residues as indicators of toxicity. 

5-18
Although many toxicological studies report measurements of metal residues in multiple tissues 

along with adverse effects, these tissue residue values may not be appropriate for use as a CBR 

threshold because metal concentrations in some tissues may have little or no relationship with 

toxicity. Furthermore, risk assessors are cautioned against extrapolating CBRs across differing 

exposure routes (food vs. water), durations, tissues, or species, because the potency of metal 

residues often differs depending on these factors. 

5.3. RISK CHARACTERIZATION 

As described in Section 2.4, Risk Characterization is the final phase of the risk 

assessment and is the culmination of the Planning, Problem Formulation, and Analysis of 

predicted or observed adverse effects. Risk Characterization produces a detailed description of 

the risk estimate(s), evaluates and summarizes the lines of evidence that support or refute the risk 

estimate(s), describes the uncertainties, assumptions and qualifiers in the risk assessment, and 

reports the conclusions of the assessment to risk managers (U.S. EPA, 2000c, 1998a). 

While there are no metal-specific methods in the Risk Characterization, there are aspects 

that are important to metals risk assessments. For example, considering multiple lines of 

evidence such as results from in situ toxicity testing or biological assessments can be valuable 

for supporting the conclusions of a risk assessment. Care should be taken, however, to evaluate 

and present the limitations associated with each line of evidence, as discrepancies may not 

always indicate underlying differences; rather, they may reflect inherent limitations of each of 

the methods. For example, biological assessment methods may not be sufficiently sensitive to 

detect the level of effects or exposures that are of concern in the risk assessment. Documenting 

assumptions and uncertainties (e.g., use of background metal concentrations rather than added 

metal or specific metal species) becomes increasingly important the closer hazard thresholds are 

to background concentrations. Risk assessors also should document all assumptions and 

uncertainties in the methods used, such as how metals bioavailability was addressed. Because 

data may not be adequate to conduct a quantitative uncertainty analysis, risk assessors should 

describe the sensitivity of the risk assessment results to key assumptions and the direction of bias 

introduced by these assumptions (i.e., under- or overestimation of risks). This is particularly 

important for national or regional assessments, where results often are intentionally based on 

organisms and conditions that enhance exposure, bioavailability, and toxicological sensitivity. 

For essential metals, risk assessors should describe the relationship of the risk threshold to 

nutritionally required levels for the organisms of concern. Risk assessment results that fall 

below nutritionally required levels are an indication that some methods or assumptions require 

additional refinement. Risk assessors should carefully document the form(s) of metals used in 

the exposure and effects assessment, as they frequently differ due to data limitations. Additional 

issues and questions that should be addressed in the risk assessment are listed in Section 2.4. 

5-19
6. TERRESTRIAL ECOLOGICAL RISK ASSESSMENT FOR METALS 

This section of the Framework provides an overview of how the principles for metals risk 

assessment apply to ecological risk assessments for terrestrial environments. Receptors typically 

considered in these assessments include soil invertebrates, plants, and wildlife species. Some 

assessments also examine effects on microbiota and soil processes. This section of the 

Framework builds on the information presented in Chapter 2 that lays out issues to be considered 

during Problem Formulation and that describes metal chemistry associated with soil systems. 

That information is not repeated here and the reader should refer to Chapter 2 for this 

information. 

6.1. METALS PRINCIPLES 

Metals have specific environmental and biotic attributes that should be considered in all 

risk assessments. These principles for metals risk assessment (see Chapters 1 and 2) apply in 

various ways to ecological risk assessments depending on the scale of the assessment (site 

specific, regional, or national). This section describes applications of the principles to terrestrial 

ecological assessments within the standard risk assessment framework. Specifically, they fall 

into the risk assessment paradigm as follows: 

Background levels Exposure Assessment 

Mixtures Exposure and Effects Assessment 

Essentiality Effects Assessment 

Forms of metals Exposure and Effects Assessment 

Toxicokinetics/toxicodynamics Exposure Assessment (bioavailability) and 

Effects Assessment (ADME and toxicity) 

6.2. CHARACTERIZATION OF EXPOSURE 

Metal exposure assessment includes characterization of the exposure routes and pathways 

specific to metals, the phase associations and chemical forms of the metals, and the expression of 

exposure and target doses in a manner consistent with defining hazard thresholds for particular 

organisms. 

6.2.1. Natural Occurrence of Metals 

At a national level, metal concentrations vary naturally in soils across the U.S. These 

variations pose challenges for conducting national assessments of risk to terrestrial ecological 

receptors. The assessor may decide to use a single toxicity level regardless of background 

concentrations for a screening type assessment (see text box on ecological soil screening levels 

or Eco SSLs) (U.S. EPA, 2003c) or may prefer to divide the country into regions of similar metal 

6-1
background levels (metalloregions). Exposure assessments should consider metal levels 

inclusive of background. 

At the regional and local (site) scales, risk 

assessors should account for the natural 

occurrence of metals either at the beginning of an 

assessment (i.e., during Problem Formulation), 

during the assessment, or when making risk 

management decisions about the implications of 

the predicted or observed levels of metals in soil

Because the national soil survey4

 is over 20 year

old, risk assessors should consider the feasibility 

of generating site-specific concentrations for 

local risk assessments. 

s. 

s 

More appropriately, risk assessors should avoid single-result assessments for the entire 

country. Rather, such assessments should be subdivided into metal-related ecoregions known as 

“metalloregions” (McLaughlin and Smolders, 2001) so that protection levels, mitigation goals, 

and ranking results will be appropriate for the suite of species naturally present within each type 

of controlling environment. This is directly analogous to the use of ecoregions when 

establishing water quality criteria (Griffith et al., 1999). The use of metalloregions provides the 

ability to account for the broad regional parameters affecting metal availability in soils and 

waters as well as for the differences in organism response to added metal. 

The metalloregion concept (McLaughlin and Smolders, 2001), although intuitively 

appropriate, has not yet been fully developed for the U.S. The country has been divided into 

ecoregions for both aquatic and terrestrial systems (Bailey et al., 1994; Bailey, 1983). These are 

based on climactic and vegetation factors and form the basis of metalloregions. EPA is still 

working to complete ecoregion maps at much finer scales for each state (see EPA Web site at 

http://www.epa.gov/wed/pages/ecoregions/ecoregions.htm). To complete the metalloregion 

concept, soil properties that affect bioavailability (e.g., pH, cation exchange capacity [CEC], and 

organic matter [OM]) should be overlaid on the ecoregions, along with soil type (e.g., sandy 

loam, clay loam) and background concentrations of metals. Similar information is needed for 

water bodies. Although this type of information is fairly current and available, soil data have not 

been updated since the mid-1970s, which may limit their usefulness. Nevertheless, work is 

under way to develop metalloregions (e.g., McLaughlin and Smolders, 2001), although it is 

likely to be several years from the time of this writing before they are available for use in a 

decision-making capacity. 

4 Schacklette, HT; Boerngen, JG. (1984). Element concentrations in soils and surficial materials of the conterminous 

United States. U.S. Geological Survey Professional Paper 1270. 105 pp. 

EPA’s Ecological Soil Screening Levels 

EPA’s Ecological Soil Screening Levels 

(Eco SSLs) for metals are national-level 

concentrations of metals in soils that are 

protective of wildlife, plants, and soil 

organisms. These values are lower than 

naturally occurring levels in some parts of 

the country. Exceedences of such levels 

does not mean that a risk exists but does 

mean that a more regional or site-specific 

assessment may be needed. 

6-2
6.2.2. Forms of Metals 

The physical and chemical forms of metals influence exposure and subsequent effects 

and can be influenced by physical/chemical conditions in the environment. National level 

assessments involve a broad range of environmental conditions and so the risk assessor should 

account for different metal species in different locations and soil types. As assessments 

transition from national, to regional, to local, the assessor should incorporate site-specific soil 

parameters that influence metal speciation (e.g., pH, CEC, clay content). National values (e.g., 

geometric mean values) of these parameters should be used, with the same recommendation as 

discussed in Section 5.1.3 on Natural Occurrence of Metals. 

6.2.3. Exposure Routes 

The major metal exposure route that the risk assessor should consider for wildlife is 

ingestion, with a minor (and often unknown) inhalation component. For plants, root uptake is 

the most important with leaf exposures secondary, with the exception of Hg where the majority 

is accumulated via foliar uptake; Cd and sometimes Pb also may be accumulated through foliage 

but amounts relative to soil exposure will vary depending upon soil conditions (e.g., pH). Plants 

may also lose metals through foliar leaching during precipitation events although to a 

significantly lesser extent than for other micronutrients such as potassium. Soil invertebrates are 

assumed to be exposed through direct contact. Pathways describe transport of the contaminant 

in the environment and include uptake and bioconcentration (e.g., dietary ingestion of a soil 

contaminant that has been taken up by plants). Principles of metal transport and fate in soils are 

applicable to assessments of risk to all terrestrial organisms and will be discussed first. 

However, because of significant differences in exposure routes and pathways between 

invertebrates, plants, and wildlife, it is more convenient to discuss exposure assessment methods 

by receptor group. 

6.2.4. Soil Transport and Fate Models Pathway of Exposure for 

Risk assessors routinely use transport and fate Terrestrial Organisms 

models (i.e., a computational model) to describe and Pathways of exposure for

quantify exposure pathways. Models are also useful in terrestrial organisms to metals 

include movement from soils through 

situations where risk assessors are trying to estimate the food web, and to a lesser extent, 

air deposition either into soils or exposure levels that are expected to result from the directly onto terrestrial receptors (e.g., 

implementation of some permitting action or remediation plants). 

measures at local, regional, or national scales. Numerous 

models are available for use; most are based on the same fundamental principles: metals are 

ubiquitous in the environment and within each media compartment they are present in 

association with water (freely dissolved metal or as organic and inorganic metal complexes), 

6-3
particles (sorbed, precipitated, or incorporated within a mineral phase), and air. The risk assessor 

can find a more detailed discussion of these processes in Section 3.2 on Fate and Transport. 

Currently, there is no single model available that encompasses all the desirable metal-specific 

features for terrestrial systems. Discussions of the available terrestrial transport and fate models, 

as well as a number of chemical equilibrium models, may be found in Allen (2002). 

6.2.5. Toxicokinetics/Toxicodynamics 

Target organ exposure levels and subsequent effects depend on how environmental 

conditions affect speciation of a metal (e.g., whether an organism actively takes up or excludes 

metals in soils and how an organism processes metals internally). See Section 3 for details on 

environmental chemistry and issues relating to bioaccumulation. Risk assessors should 

specifically address bioavailability and bioaccumulation for each metal of concern in each 

environment (either a local site for site-specific assessments or some larger estimate for regional 

and national level assessments). 

6.2.5.1. Bioavailability 

Risk assessors should adjust bulk soil metal concentrations by appropriate bioavailability 

factors to achieve comparable, actual uptake of metals by soil organisms. This will standardize 

exposure values across soil types and allow for more accurate comparisons with laboratory 

toxicity data. Cation exchange capacity (CEC) recently has been shown to be an important 

factor modifying zinc bioavailability in soils, and presumably it will be important for other 

cationic metals as well. However, CEC is strongly dependent on the type and amount of organic 

material (OM) and oxyhydroxides present in the soil, and is strongly pH dependent. Surface 

charge on OM and oxyhydroxides increases with pH, thereby increasing their sorptive capacity 

for metals (thus decreasing metal bioavailability). Conversely, positive surface charges increase 

as the pH drops, which increases sorption of anions (e.g., As or Se) under low pH conditions and 

decreasing sorption of cation ionic metals. Clays, on the other hand (except for kaolinite), have a 

surface charge that is largely independent of pH. Therefore, normalization of toxicity data to 

CEC can be done only within specific soil types and pH ranges, which frequently are not 

specified either in laboratory bioassays or many field studies. Furthermore, it is important for 

the risk assessor to note that most published values of CEC are measured at pH 7. In general, 

risk assessors can assume that cationic metals are more bioavailable at lower soil pH (<6) and 

less bioavailable at higher soil pH (>8). The opposite assumption holds for anionic metals. 

Soil chemical models are being developed to predict how aging will modify bulk soil 

concentrations when soils are amended with soluble salts. Aging reduces the bioavailable 

fraction of metals over time. Preliminary studies suggest that consideration of aging may result 

in estimates of the bioavailable fraction as low as 0.1 × bulk soil concentrations (McLaughlin et 

6-4
al., 2002). Until the data become available for metals of concern, toxicity values derived from 

soluble-salt amended soils (which have not simulated aging) cannot be reliably corrected to 

approximate aged metals in field situations and the risk assessor should acknowledge this as a 

significant uncertainty during the risk characterization. 

Ideally, exposure should be expressed on the basis of pore water concentration, to 

account for all factors influencing bioavailabilty; however, there are currently significant 

limitations to collecting and interpreting metal-related data from soil pore waters and such 

information generally is not available (even at site-specific assessments and never for regional or 

national assessments). The risk assessor could estimate metal concentration in soil pore water 

using EqP theory (as with sediment pore water analyses; see Section 3.1.5). The risk assessor 

can use published soil binding coefficients (Kds) to estimate partitioning between soil particles 

and pore water although these values also are inherently uncertain (published value depends on 

derivation method, soil type, etc.). Furthermore, toxicity threshold values generally are provided 

as bulk soil concentrations so the risk assessor would not be able to compare pore water exposure 

with any effects estimates. 

6.2.5.2. Bioaccumulation 

For terrestrial ecosystems, the concept of bioaccumulation is intended to capture the 

potential for two ecologically important outcomes: (1) direct toxicity to plants and wildlife and 

(2) secondary toxicity to animals feeding on contaminated plants and animals. This approach 

stresses the potential for trophic transfer of metals through the food web, so total exposure can be 

calculated, including dietary intake as well as intake from contaminated environmental media 

(soil and water). For vegetation or soil invertebrates, the bioaccumulation factor (BAF; or biotasoil accumulation factor, BSAF) is defined as field measurements of metal concentration in plant 

tissues divided by metal concentration in soil (or soil solution); the BCF is defined as the same 

measurement carried out in the laboratory (Smolders et al., 2003). 

Risk assessors should be aware that data applicability is directly related to which tissue is 

sampled and how it is processed. BAFs for plants include metals aerially deposited on leaves as 

well as those in soil particles adhering to roots. Such metals will not be part of BCFs, which 

frequently are determined in hydroponic culture. Similar differences between BCFs and BAFs 

apply for earthworms exposed in soils versus laboratory studies using the filter paper substrate 

protocols. Furthermore, BCFs within earthworms may not include additional feeding of the 

animals during the study. Field studies are reflective of chronic exposures, whereas BCFs may 

be calculated from shorter time frames. Ideally, risk assessors should select BCFs reported at 

equilibrium (i.e., after sufficient exposure time to maximize the BCF). Whole-body BAFs 

generally are not calculated for birds and mammals, except for small mammals such as rodents 

(Sample et al., 1998b). Risk assessors should understand the conditions under which metal 

concentrations were measured and critically examine data to determine whether they are reported 

6-5
as wet or dry weight (the ratio of tissue to soil concentrations must be done on the same wet/dry 

weight basis for both). 

For soil invertebrates and most plants, metal BAFs are typically less than 1 and usually 

are based on the total metal in soil and tissue that do not account for bioavailability differences. 

The risk assessor might consider using a ratio of total metal in the organism to some measure of 

the bioavailable fraction of metal in the soil (e.g., free ion concentration or weak salt extractable) 

for expressing a BAF to allow comparison among different soils, although, in general, data are 

lacking for using this method. 

Furthermore, the risk assessor is reminded that bioaccumulation of metals is not a simple 

linear relationship. Uptake is nonlinear, increasing at a decreasing rate as medium concentration 

increases. Models for predicting metal bioaccumulation by soil invertebrates are primarily 

statistical in nature, describing relationships between metal body burdens in oligochaetes and 

collembola, soil metal concentrations, and soil physical/chemical characteristics. Sample et al. 

(1998a) and Peijnenburg et al. (1999b) have each developed univariate uptake models for 

earthworms that are based on empirical data (metal concentrations in worms vs. the natural log 

of amount of metal in soils) that risk assessors can use as a first approximation for 

bioaccumulation in soil invertebrates; however, these models are not specific to soil type and, 

therefore, do not account for bioavailability factors. Furthermore, they do not adequately predict 

Cr or Ni uptake. An alternative approach that the risk assessor could consider is the use of 

multivariate statistical models to look for patterns of uptake of multiple metals to predict the 

potential bioconcentration of one metal of particular interest (Scott-Fordsmand and Odegard, 

2002) or BAF as a function of soil characteristics (Saxe et al., 2001; Peijnenburg et al., 

1999a, b). Path analysis has been suggested as an alternative for multiple regression in 

describing these relationships. It partitions simple correlations into direct and indirect effects, 

providing a numerical value for each direct and indirect effect and indicates the relative strength 

of that correlation or causal influence (Bradham, 2002; Basta et al., 1993). 

The absolute level of metal accumulation is not as important as the rate of uptake (Hook 

and Fisher, 2002; Hook, 2001; Roesijadi, 1992). Adverse effects are avoided as long as the rate 

of metal uptake does not exceed the rate at which the organism is able to bind the metal, thereby 

preventing unacceptable increases in cytosolic levels of bioreactive forms of the metal. If the 

rate of uptake is too great, the complexation capacity of the binding ligand (e.g., metallothionein) 

could be exceeded; cytosolic metal levels then become unacceptably high, and adverse effects 

can ensue. Because measures of uptake rates are not available, static concentrations are used; the 

risk assessor should acknowledge this uncertainty during the Risk Characterization. 

6.2.6. Soil Invertebrate Exposure 

The soil ecosystem includes a complex food web of soil invertebrates (both hard- and 

soft-bodied) that feed on each other, decaying plant material, and bacteria or fungi. However, 

6-6
the risk assessor should estimate exposure as a function of soil concentration, rather than as a 

detailed analysis of movement of metals through the food web, to generate data that will be 

comparable to effects concentrations. This is a reasonable approximation for soft-bodied 

invertebrates (e.g., earthworms) whose metal exposure is primarily through soil pore water (from 

both dermal absorption and soil ingestion) (Allen, 2002). There is more uncertainty in 

correlating soil metal concentrations with effects in hard-bodied invertebrates because they are 

primarily exposed through ingestion of food and incidental amounts of soil (Sample and Arenal, 

2001). Regardless, risk assessors should estimate soil invertebrate exposure on the basis of total 

metal concentration in bulk soils (adjusted for relative bioavailability, where possible) collected 

in the top 0-12 cm of soil (U.S. EPA, 2003c, 1989b). In detailed, site-specific assessments, the 

organic matter on top of the soil (the “duff”) may be analyzed separately to provide further detail 

on exposure to detritivores (such as Collembola) and deeper-soil-dwelling organisms (e.g., 

various species of earthworms). 

6.2.7. Plant Exposure 

Plants access metals through the pore water although mycorrhyzae, protons, and 

phytosiderophores released by the root can significantly influence the microenvironment and 

change uptake rates of metals (George et al., 1994; Sharma et al., 1994; Laurie and Manthey, 

1994; Arnold and Kapustka, 1993). Furthermore, plants have both active and passive 

mechanisms for taking up or excluding metals, depending on internal concentrations and whether 

or not the metal is an essential micronutrient, or whether it is mistaken for an essential 

micronutrient. Plants can be exposed to metals via aerial deposition onto leaf surfaces, trapping 

metals in hairs or rough cuticular surfaces. This might provide an exposure route for herbivores; 

it may also provide an exposure route for plants, as there are ion channels through the cuticle that 

are able to transport ionic metals from the leaf surface to other locations in the plant, depending 

on the inherent mobility of the metal in the xylem and phloem (Marschner, 1995). 

The risk assessor should consider the default approach to estimating exposure of plants to 

metal as measuring metal concentrations in bulk soil (top 0-12 cm). However, as with soil 

invertebrates, this overestimates exposure because it does not account for differential 

bioavailability and aging. The risk assessor generally can categorize metal bioavailability and 

uptake based on soil pH and organic matter (see Section 3.1.6.5). It is very clear that strongly 

acidic soils increase plant uptake of Zn, Cd, Ni, Mn, and Co and increase the potential for 

phytotoxicity from Cu, Zn, and Ni. Alkaline soil pH increases uptake of Mo and Se, while Pb 

and Cr are not absorbed to any significant extent at any pH (Chaney and Ryan, 1993). 

6-7
Table 6-1. Qualitative bioavailability of metal cations in natural soils to 

plants and soil invertebrates 

Soil type 

Soil pH 

Low 

organic matter 

(<2%) 

Medium 

organic matter 

(2 <6%) 

High 

organic matter 

(6 to 10%) 

4 # Soil pH # 5.5 Very high High Medium 

5.5 < Soil pH # 7 High Medium Low 

7 # Soil pH # 8.5 Medium Low Very low 

Table 6-2. Qualitative bioavailability of metal anions in natural soils to 

plants and soil invertebrates 

Soil type 

Soil pH 

Low 

organic matter 

(< 2%) 

Medium 

organic matter 

(2 to <6% ) 

High 

organic matter 

(6 to 10%) 

4 # Soil pH #5.5 Medium Low Very low 

5.5 < Soil pH < 7 High Medium Low 

7 # Soil pH # 8.5 Very high High Medium 

Source: U.S. EPA (2003c). 

Qualitative relationships between soil chemistry and bioavailability are appropriate for 

national-scale application. However, for site-specific or metals-specific applications, the risk 

assessor should use quantitative methods. Parker and Pedler (1997) and Lund (1990) have 

suggested that only uncomplexed, free ionic species of cations can be taken up by roots, and this 

has been described using a Free Ion Activity Model (FIAM) similar to the Biotic Ligand Model 

(BLM) used in aquatic systems. However, significant exceptions to the free-ion model have 

been identified; so until this theory is tested more thoroughly, the risk assessor should continue 

to estimate exposures using bulk soil values with qualitative estimates of bioavailability based on 

soil type (pH and OM). Again, the risk assessor should acknowledge these uncertainties during 

the Risk Characterization. 

6.2.8. Wildlife Exposure 

The relative importance of exposure pathways and routes varies by animal species and by 

metal, although, in general, wildlife exposure is primarily through diet and incidental ingestion 

6-8
of soils or sediments. There are certain chemicals and 

exposure situations for which inhalation or dermal pathways 

are important, but in most situations the risk assessor can 

consider them to be insignificant contributors to total metal 

loads (U.S. EPA, 2003c). 

Wildlife food chain exposures for metals are 

controlled by bioavailability, bioaccessibility, and 

bioaccumulation. Bioaccessibility of metals to animals and plants that live on or in the soils can 

be influenced by soil parameters, such as pH, CEC, and organic carbon. These soil parameters 

tend to be less important for soils that are incidentally ingested by wildlife species. 

The relative importance of exposure pathways (soil vs. food chain) is dictated by the 

fraction of metal-contaminated soil in the diet and the amount of accumulation of metal in food 

items. In the absence of site-specific information, the risk assessor can use the following 

generalizations to determine the relative importance of incidental soil ingestion versus dietary 

metals: 

Exposure Pathway for 

Terrestrial Wildlife 

Food and the incidental 

ingestion of soil are the two 

most important exposure 

pathways for terrestrial wildlife. 

1. Incidental soil ingestion is a proportionally more important pathway for herbivores 

than for carnivores or invertivores. 

2. Uptake into soil invertebrates (e.g., earthworms) is a proportionally more important 

pathway for animals that feed on these organisms. (Note: This assessment reflects 

work done with earthworms and may not apply to hard-bodied soil invertebrates such 

as Colembolla.) 

3. If bioaccumulation is low (<<1), importance of soil ingestion versus diet for metal 

exposure increases. 

4. When bioaccumulation is greater (~1 or higher), the food pathway should dominate. 

5. The closer the association an animal has to the ground, the greater the importance of 

soil ingestion. This association may be due to ground foraging, burrowing habits, etc. 

6. The looser the association with the ground (e.g., piscivores, aerial/arboreal 

insectivores, raptors), the lower the importance of soil ingestion. 

Figure 6-1 provides a simple scheme for the risk assessor to use for judging the relative 

contribution of food and soil before accounting for bioavailability. The assessor should assume 

that incidental ingestion of soil becomes proportionally more important for exposure to wildlife 

when (1) the bioaccumulation factor (BAF) from soil to food (e.g., to plants or soil invertebrates) 

is less than 1 and (2) the fraction of soil in the diet is greater than 1%. However, the risk 

assessor should use these generalizations with caution for site-specific assessments. As the risk 

assessor acquires more site-specific information, the relative importance of pathways may 

6-9
change. For example, site-specific data may show that the accumulation of a chemical into 

plants or soil invertebrates is much lower than indicated by the default assumptions. In such 

cases, incidental ingestion of soil would become proportionally more important. The 

bioavailability of metals in incidentally ingested soil is also variable. Therefore, when the 

exposure is being driven by incidental soil ingestion, the risk assessor should consider 

refinements of exposure estimates through a better understanding of bioavailability, although 

very little information is available on this for most wildlife species. 

100 

80 

60 

40 

20 

0 

Increasing Percent Soil in Diet 

BAF = 0.01 

BAF = 0.1 

BAF = 1 

BAF - 10 

Percent contribution of soil 

ingestion to total dose 

Figure 6-1. Generalized representation of percent contribution of incidental 

soil ingestion to oral dose for wildlife at different soil ingestion rates and 

bioaccumulation factors and a bioavailability of 100 percent. 

The risk assessor should be cautious about extrapolating bioavailability adjustments for 

wildlife from models developed for estimation of bioavailability of metals in soils for incidental 

human exposures. There are significant variations in digestive physiology and anatomy across 

mammalian and avian species that alter the degree of assimilation and uptake of metals (MenzieCura and TN&A, 2000). For example, metals present in soils may be more or less bioavailable 

within the gut of an herbivore that relies on fermentation as compared to the simpler gut of a 

carnivore that is designed to break down proteins. These gut systems differ in chemistry 

(including pH) and residence time. 

Food chain modeling can be used to estimate the exposure of wildlife to metals based on 

the ingestion of soil, food, and water. The risk assessor should use the same dietary uptake 

6-10
model for metals as is used for organic substances, e.g., Eco SSLs; Ecological Committee on 

FIFRA Risk Assessment Methods (ECOFRAM) (Sample et al., 1997; U.S. EPA, 1997d). For 

national or regional risk assessments, the assessor may use trophic transfer rates to model food 

concentrations but only on the basis of soil measurements (rather than using direct measures of 

concentration of metals in food items). As with aquatic organisms, trophic transfer values for 

metals in terrestrial systems are an inverse function of soil concentrations. Therefore, the risk 

assessor should not use constants for this term but rather should generate regression equations of 

plant and invertebrate uptake rates as a function of soil concentrations and use which ever 

value(s) that are consistent with the degree of conservatism or amount of realism appropriate for 

the assessment. Sample et al. (1998a) developed uptake models to predict concentrations in 

earthworms from soil concentrations and Efroymson et al. (2001) provides similar information 

for plants (see Section 6.5.2.3 Bioaccumulation for a more detailed discussion). 

With the exception of a few hyperaccumulator species, the risk assessor can reasonably 

assume that most plant species do not bioconcentrate metals (i.e., BAFs <1). Pb, As, Cr, and Co 

are not taken up by plants in measurable quantities, and the small amount that is taken up is 

mostly confined to root tissues (Chaney et al., 2000; McGrath, 1995; Chaney and Ryan, 1994; 

Xu and Thornton, 1985). In contrast, many plants are quite sensitive to some metals (Mn, Zn, 

Cu, for example); the risk assessor should be aware that plants frequently die before achieving 

high metal concentration levels that pose a threat to animals via food chain transfer (with the 

exception of the hyperaccumulator species, as noted above). 

6.3. CHARACTERIZATION OF EFFECTS 

When assessing metal toxicity to terrestrial organisms, the risk assessor should 

understand both the natural mechanisms of tolerance for (or, in the case of micronutrients, the 

use of) metals and the toxicological responses that occur when exposure exceeds the capacity of 

the organism to regulate its body burdens. The risk assessor should also consider interactions 

between metals in either their uptake or toxicity (such as Cd/Ca/Zn, Hg/Se, Cu/Mo). Risk 

assessments for metals are further complicated by the need to express the dose-response (or 

concentration-response) functions in bioavailable units that are functionally equivalent to 

measures of exposure. This section provides tools and approaches risk assessors can use when 

addressing issues of essentiality, metal mixtures, 

and appropriate use of toxicity tests; issues of how 

acclimation or adaptation to continued exposures 

may affect toxicity have been addressed in Sections 

1.4.1 and 4.2.1 on Natural Occurrence of Metals. 

Essentiality 

Essentiality, or the requirement for 

normal organism metabolic function, of 

many metals is one of the primary factors 

that differentiates risk assessment for 

metals and metal compounds from that of 

synthetic organic chemicals. 6.3.1. Essentiality 

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Essentiality, or the requirement for normal organism metabolic function, of some metals 

is one of the primary factors that differentiates risk assessment for metals and metal compounds 

from that of synthetic organic chemicals (Janssen and Muyssen, 2001). Some trace elements, 

such as Co, Cu, Fe, Mn, Se, Mo, and Zn, are necessary for the normal development of plants and 

animals. Other metals, such as As, Cd, Pb, and Hg, have no known functions in plants and 

animals (Mertz, 1981). Table 6-3 classifies the metals addressed in this Framework by their 

known essentiality to organisms. 

The risk assessor should be sure that effects thresholds such as Toxicity Reference 

Values (TRVs) are not lower than the nutritional requirements for the particular plant or animal 

species being evaluated. If TRVs are set too 

low (i.e., in the range where deficiency can 

occur), the determination of risk will be 

erroneous and deficiency effects will be 

mistaken for toxic responses. For wildlife, the 

risk assessor can consult the literature on dietary 

requirements of essential elements for livestock 

(McDowell, 2003; NAS/NRC, 1994a, 1980). 

Marschner (1995) summarizes the minimum 

concentrations required for plant growth. 

Because of differences in test conditions 

among published studies, it may be difficult for 

the risk assessor to directly compare toxicity 

threshold values with recommended dietary 

requirements of essential elements. 

Extrapolation of data among species (e.g., from livestock to wildlife species) may also add 

uncertainty to the effects assessment. Furthermore, addition of safety factors when deriving 

protective values often results in concentrations significantly below required intake. The risk 

assessor should address these and similar uncertainties in toxicity threshold derivations as part of 

the Risk Characterization process. Detailed site-specific assessments, where more accurate 

estimates of effects thresholds are expected, may require the risk assessor to request additional 

bioassays to characterize the biphasic dose-response curve and determine both required and 

excessive threshold levels. 

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Threshold Values 

For essential elements, it is important to 

ensure that effects thresholds, such as 

Toxicity Reference Values (TRVs), are not 

lower than the nutritional requirements for 

the plant or animal species being evaluated. 

It may be difficult, however, for the risk 

assessor to directly compare toxicity 

threshold values with recommended dietary 

requirements because of differences in test 

conditions among published studies. 

In screening-level assessments, toxicity 

threshold values can be used by the risk 

assessor, if they are not lower than 

estimated requirements. Detailed, higher 

level assessments may require additional 

bioassays to characterize the biphasic doseresponse curve and determination of both 

required and excessive threshold levels. 
Table 6-3. Metals classified by their known essentiality 

Metal 

Essential 

(known requirement for 

health and function) 

Beneficial 

(but not known to be 

essential) 

Nonessential 

(and not known to 

be beneficial) Plants Animals Plants Animals 

Aluminum (Al) x 

Antimony (Sb) x 

Arsenic (As) X 

Barium (Ba) x 

Beryllium (Be) x 

Cadmium (Cd) x 

Chromium (Cr) x 

Cobalt (Co) x x 

Copper (Cu) X x 

Lead (Pb) x 

Manganese (Mn) X x 

Mercury (Hg) x 

Molybdenum (Mo) X x 

Nickel (Ni) X x 

Selenium (Se) x x 

Silver (Ag) X 

Strontium (Sr) X 

Thallium (Tl) X 

Vanadium (V) X 

Zinc (Zn) X x 

Source: Adapted from a table presented in SRWG (2002) and incorporating data from NAS/NRC (1980) 

and Barak (1999). Fairbrother and Kapustka (1997) discussed the roots of essentiality of naturally 

occurring elements. 

6.3.2. Toxicity Tests 

For assessments conducted for regional or national assessments, criteria development, or 

ranking purposes, risk assessors should acknowledge that results will be based on organisms and 

soil types that result in greatest bioavailability and sensitivity. The risk assessor should take 

great care to ensure that the organism-environment combinations that are assessed are, in fact, 

6-13
compatible with real-world conditions. Thus, for site-specific assessments, species tested and 

water (or sediment) used in the test system should be similar to conditions at the site. In the 

absence of such information, risk assessors could use data from standard test species and 

conditions, but uncertainty factors may be warranted to adjust the final toxicity value. 

6.3.3. Metal Mixtures 

Mixtures of metals (including metalloids and other contaminants) are commonly 

encountered in the natural environment as a result of anthropogenic inputs and should be 

considered by the risk assessor for all assessments. Metal interactions, according to Calamari 

and Alabaster (1980), occur at three levels: 

1. Chemical interactions with other constituents in the media, 

2. Interactions with the physiological processes of the organism, and 

3. Interactions at the site of toxic action. 

The joint action of metal mixtures may be expressed in different ways, such as increasing 

or decreasing the toxicity relative to that predicted for individual components. As a result, the 

toxicity of metal mixtures has important consequences for metals risk assessments. However, 

predicting the toxicity of metal mixtures has proven to be a difficult challenge in ecotoxicology. 

Much of the difficulty in interpreting the available information on the toxic effects of 

metal mixtures is due to differences in the bioavailability of metals (and measures used to define 

the bioavailable fraction) that occur across mixture studies. As discussed in Section 3, the 

bioavailability of metals depends on a suite of factors affecting their speciation, complexation 

with ligands, and interaction with biological systems. Nevertheless, the risk assessor needs some 

measure of the bioavailable metal fraction in the exposure media to accurately predict the effects 

of metals mixtures (Sauvé et al., 1998; Weltje, 1998; Posthuma et al., 1997). Besides 

bioavailability issues, the joint action of metal mixtures can depend on the overall mixture 

concentrations and the relative proportion of the constituent metals, as has been seen in aquatic 

studies (Norwood et al., 2003; Mowat and Bundy, 2002; Fargašová, 2001; Sharma et al. 1999). 

The two most common classes of models used to predict mixture toxicity are the 

Concentration Addition and Effects Addition models. These models have been used to classify 

the combined effects of chemical mixtures as being less than additive (i.e., when the observed 

effect is less than the model prediction), strictly additive (i.e., matching model predictions), and 

more than additive (i.e., when the observed effect is greater than model predictions; Norwood et 

al., 2003). Both models use metal concentrations in media to generate concentration-response 

curves for individual metals, and these data are then used to generate specific critical 

concentrations for mixture models. In the Concentration Addition model, all metals in a mixture 

are added together to predict toxicity; differing potencies are taken into account by converting 

6-14
chemical concentrations to an equitoxic dose (e.g., Toxic Units (TUs) or Toxicity Equivalence 

Factors (TEFs), which converts all metals to one metal concentration). Concentration Addition 

is used often when the constituents are known or assumed to act through the same or similar 

MOA. However, the risk assessor should use caution when applying the Concentration Addition 

model to mixtures containing many metal constituents (particularly those well below toxic 

levels) because of the potential for an upward bias in predicted mixture toxicity (Newman et al., 

2004). In the Effects Addition model, differing potencies are ignored, and the effect of each 

metal’s concentration is combined to predict mixture toxicity. The Effect Addition model is 

often used when constituents act independently (i.e., different modes of action). Only the 

Concentration Addition model allows detection of toxicity that is more than additive. Thus, a 

key issue in applying either the Concentration Addition or Effects Addition model is to define 

the nature of the metals’ joint action (i.e., independent or similar mode of action). The risk 

assessor can use information on the MOA, capacities to act as analogues for other metals, 

essentialities and ligand binding tendencies to make this decision. 

Risk assessors should keep in mind, however, that toxicities of certain metal elements are 

associated with deficiencies of others. For example, increased Zn, Cu, and Ni toxicities can be 

associated with Fe deficiencies (Bingham et al., 1986), and increased Pb and Zn toxicities can 

also be related to P deficiencies (Brown et al., 2000, 1999; Laperche et al., 1997). The behavior 

of plant species in response to nutrient deficiencies varies, and this behavior can affect the uptake 

of metal elements (Marschner, 1998). Similar interactions occur in wildlife; for example, Cu 

toxicity can be a result of Modeficiency and vice versa (McDowell, 2003; NAS/NRC, 1994a, 

1980). 

It is possible that receptor binding models (e.g., FIAM) may be expanded in the future to 

include mixtures. In theory, if two metals compete for binding to the same site of toxic action on 

an organism, it should be possible to model the total metal bound to that site and, hence, to 

predict metal toxicity using a mechanistic receptor binding approach in an Effects Addition 

model. Alternatively, if two metals do not compete for the same binding site on the organism, 

then these models may provide more reliable estimates of individual metal bioavailability, and 

these estimates can then be combined in more accurate Effects Addition models. However, at 

present, these possibilities remain theoretical. Additionally, this possibility, while improving the 

ability to assess the effects of metal mixtures, does not include temporal aspects (i.e., “time-toresponse” versus concentration). 

From the preceding discussion, it should be clear that the accurate prediction of joint 

toxicity of metal mixtures to terrestrial organisms remains a significant challenge. 

6.3.4. Critical Body Residues 

Critical body residues (CBRs) are internal concentrations of chemicals that are correlated 

with the onset of a toxic response (Conder et al., 2002; Lanno et al., 1998). CBRs can be based 

6-15
on whole-body residues (see below for discussion of this approach in soil invertebrates) or 

concentrations in specific tissues. The risk assessor may choose to use CBRs instead of dietary 

TRVs to reduce uncertainties because they account for site-specific bioavailability and 

multipathway issues (Van Straalen, 1996; Van Wensem et al., 1994). Unfortunately, there are 

major data gaps in available CBRs for many species—metal combinations. 

Risk assessors can use tissue-specific critical loads for some metals that have been 

established for several species of vertebrate wildlife, including Pb in liver, Cd in kidney, Hg in 

brains, and Se in eggs. See Beyer et al. (1996) for these figures. Only a few CBRs have been 

developed in soil invertebrates for metals (Conder et al., 2002; Crommentuijn et al., 1997, 1994; 

and Smit, 1997 for Cd and Zn). 

For plants, the use of a tissue residue (CBR) approach is another method that risk 

assessors might use to address metal toxicity issues, based on the concept that a metal 

concentration must reach a threshold value in the organism or at the target site before effects 

begin to occur (McCarty and Mackay, 1993; Lanno and McCarty, 1997). For essential elements 

in plants, deficiency/sufficiency concentrations in foliage have been developed. However, the 

relationship between toxicity and tissue residues is complex and varies depending on tissue type 

(roots vs. shoots), plant species, and metal and there is little to no information available. 

Therefore, this approach, although conceptually sound, requires significant research before risk 

assessors will find it useful. 

6.3.5. Plant and Soil Invertebrate Toxicity 

The risk assessor can estimate TRVs (i.e., toxic thresholds) for plants and soil 

invertebrates from laboratory tests where metals are mixed with standard soils (Fairbrother et al., 

2002). Variability among soil toxicity test results is due in part to the influence of soil properties 

on bioavailablity of metals (e.g., pH, organic matter and CEC). Additionally, acclimation and 

adaptation of test organisms can further complicate test results and aging and other 

physical/chemical processes that affect metal speciation and uptake are not represented. Because 

incorporation of sparingly soluble substances, such as many environmental forms of metals, into 

the soil matrix is difficult, tests generally are conducted using soluble metal salts with the 

addition of organism to the test matrix immediately after mixing. The risk assessor should be 

aware of how all these factors influence the test outcome and subsequent TRV derivation. 

There is a large body of literature on toxicity of metals to soil organisms (e.g., van 

Straalen and Løkke, 1997), although often the objectives were to understand processes rather 

than to develop defensible toxicity thresholds. The challenge for the risk assessor, therefore, lies 

in how to use these data, taking into account the test-to-test variability in soil chemistry 

parameters, and how to develop a technically defensible means of extrapolating toxicity 

responses across soil type—in other words, how to adjust the toxicity threshold values for 

bioavailability differences in test conditions. One approach to addressing variability in soil 

6-16
toxicity tests is to normalize test results by dividing the LC50 (or, more generally, the ECx) by 

percent organic matter (Lock and Janssen, 2001). This approach is based on observed 

correlations between the LC50 of Cu to earthworms and soil organic matter content (Lock and 

Janssen, 2001). More recently, CEC has been shown to be the most important factor modifying 

Zn bioavailability in soils for both invertebrates and plants. Because CEC is a function, at least 

in part, of soil pH, normalization using this parameter should be done only among soils of similar 

pH ranges. However, comparison of field data with laboratory toxicity response information still 

is best accomplished by measuring metals in soil pore water from field assessments and 

comparing such data to spiked laboratory soils. Risk assessors can use the guidance document 

developed for establishing Ecological Soil Screening Levels or Eco SSL to judge the 

applicability of literature studies to plant or soil invertebrate toxicity threshold determinations. 

Eco SSLs have been developed for several metals, and the risk assessor should refer to these for 

national or regional assessments and for screening level, site-specific assessments. 

6.3.6. Wildlife Toxicity 

Toxicity in wildlife from metals exposures is generally poorly understood and is rarely 

quantified in field settings. A few notable exceptions are those mechanisms described in avian 

waterfowl exposure to Se (Adams et al., 2003), exposure of waterfowl to Pb-contaminated 

sediments (Henny et al., 2000; Beyer et al., 1998; Blus et al., 1991), and white-tailed ptarmigan 

exposure to Cd in vegetation (Larison et al., 2000). Most metals express multiorgan toxicity, 

resulting in a decrease in overall vigor, as opposed to well-defined mechanisms of action 

documented from organic xenobiotics such as pesticides. Typically, toxicological data used to 

assess the risk of many metals to wildlife are derived from laboratory species such as rats, mice, 

or domestic livestock species (e.g., cattle and chickens) exposed to soluble metal salts. Risk 

assessors will need to extrapolate the results of such tests to species of interest because of the 

paucity of data on the toxicity of metals to wildlife. However, risk assessors should approach 

this carefully due to the large amount of uncertainty that could be introduced into the risk 

assessment process (Suter, 1993). 

Laboratory and domestic species may be more or less sensitive to chemicals than are the 

selected wildlife species. Toxicological responses vary among species because of many 

physiological factors that influence the toxicokinetics (absorption, distribution, and elimination) 

and toxicodynamics (relative potency) of metals after exposure has occurred. For example, 

differences in digestive tract physiology, renal excretion rates, and egg production influence the 

toxicokinetics of metals. The ability of some species to more rapidly produce protective proteins 

such as metallothionein after exposure to metals is a toxicodymamic features leading to 

interspecific extrapolation uncertainty. Thus, risk assessors should not extrapolate data from 

mammal studies to birds, and should be aware that extrapolation of data from rats (simple, 

monogastric digestive physiology) to ruminants introduces more uncertainty than does 

6-17
extrapolation from rats to canids, and so on. 

In the case of metals, which some species 

are able to regulate or store in their tissues 

without experiencing toxic effects (i.e., 

biota-specific detoxification), extrapolations 

between species used to assess 

bioaccumulation and toxicity can be 

especially problematic. These difficulties in 

interspecific extrapolations are not unique to 

metals risk assessment except when dealing 

with essential elements. A review of 

potential extrapolation methodologies can 

be found in Kapustka et al. (2004). 

Currently, the best sources of 

information for the risk assessor on wildlife metal toxicity thresholds are NAS/NRC (1994a, 

1980), McDowell (2003), and the documentation supporting development of Eco SSLs values. 

The Eco SSL document also includes an approach for screening studies for acceptability for use 

in derivation of toxicity thresholds for risk assessments, which can then be used for deriving sitespecific TRVs for the most applicable endpoints. Risk assessors should apply uncertainty factors 

for extrapolation of data to species in a different taxonomic category (e.g., genus, family or 

class) with caution and include a discussion of uncertainty in the risk characterization. 

Summaries for some metals are available in Beyer et al. (1996) and Fairbrother et al. (1996). 

Risk Characterization 

Have the qualitative assessment, quantitative 

assessment, and key uncertainties regarding 

metals been presented in accordance with 

EPA guidelines? 

Do conclusions fully reflect risks in relation to 

ambient concentrations, essentiality of metals, 

chemical speciation, and information on 

variability in species sensitivity? 

Have assumptions and uncertainties been 

documented adequately? 

Have available data on mechanisms of action 

and metal interactions been fully explored in 

developing the quantitative assessment in 

accordance with EPA Guidance on Mixtures 

Risk Assessment? 

6.4. RISK CHARACTERIZATION 

Risk Characterization is the final phase of the risk assessment process, in which 

information from hazard characterization; dose-response assessment and exposure assessment 

are jointly considered to determine the actual likelihood of risk to exposed populations (U.S. 

EPA, 2000c, 1998a). The characterization also should discuss the uncertainties in the exposure 

and effects assessments, and the level of confidence in the overall determination of risk. At the 

same time, Risk Characterization is the first phase in the risk management process, in which 

information from the characterization is integrated into the consequences of rule-making or risk 

management, such as consideration of cost, alternative solutions, political considerations, 

community interactions. 

Each Risk Characterization should include three components: a qualitative summary of 

each section of the risk assessment, a numerical risk estimate, and a description of assumptions 

and uncertainties. These descriptions of variability and uncertainty are particularly important for 

metals risk assessments given all the components and challenges discussed in this Framework 

document. These are in addition to the variability and uncertainties that are inherent in all risk 

6-18
assessments (e.g., species to species toxicity extrapolations). Because information, knowledge, 

and tools are lacking for many of the metal-specific uncertainties, risk assessors should be 

particularly diligent in documenting whether these may result in an over- or under-estimation of 

risk (i.e., result in a conservative risk estimate or not). It is likely that site-specific risk 

assessments will have fewer uncertainties than regional or national scale assessments because 

risk assessors have access to local data on key issues such as specific metal species, relative 

bioavailability, or background metal levels. For national or regional assessments, selection of 

ranges or specific numbers for these values will depend upon the degree of conservatism desired 

by the risk assessor and, therefore, should be clearly documented during the Risk 

Characterization phase. 

6-19
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APPENDIX A: FRAMEWORK FOR METALS RISK ASSESSMENT CONTRIBUTORS 

AND REVIEWERS 

CONTRIBUTORS AND REVIEWERS 

Environmental Chemistry: 

Kim Anderson, Oregon State University, Corvallis, OR 

Rufus Chaney, U.S. Department of Agriculture, Beltsville, MD 

Paul Chrostowski, CPF Associates, Inc., Takoma Park, MD 

Miriam L. Diamond, University of Toronto, Toronto, ON, Canada 

Donald Langmuir, Hydrochem Systems Corp., Colorado School of Mines, Golden, CO 

Deborah Luecken, U.S. EPA, Office of Research and Development, National Exposure Research 

Laboratory, Research Triangle Park, NC 

Michael Newman, College of William and Mary, Gloucester Point, VA 

Bernard Vigneault, CANMET Mining & Mineral Sciences Laboratories/Natural Resources, 

Ottawa, ON, Canada 

Bioavailability and Bioaccumulation: 

Michael Beringer, U.S. EPA, Region VII, Kansas City, KS 

John Drexler, University of Colorado, Boulder, CO 

Nicholas Fisher, State University of New York, Stony Brook, NY 

Gerry Henningsen, H&H Scientific Services, Centennial, CO 

Roman Lanno, Ohio State University, Columbus, OH 

Jim McGeer, Natural Resources Canada, Ottawa, ON, Canada 

Exposure: 

Gary Diamond, Syracuse Research Corp., Syracuse, NY 

Charles A. Menzie, Menzie-Cura & Associates, Inc., Chelmsford, MA 

Jacqueline Moya, U.S. EPA, Office of Research and Development, National Center for 

Environmental Assessment, Washington, DC 

Michael Newman, The College of William and Mary, Gloucester Point, VA 

Jerome Nriagu, University of Michigan, Ann Arbor, MI 

Human Health Effects: 

Ed Bender, U.S. EPA, Office of Research and Development, Office of Science Policy, 

Washington, DC (Retired) 

Michael Beringer, U.S. EPA, Region VII, Kansas City, KS 

Harlal Choudhury, U.S. EPA, Office of Research and Development, National Center for 

Environmental Assessment, Cincinnati, OH 

Dominic Di Toro, University of Delaware, Newark, DE 

Mari Hirtzel Golub, University of California, Davis, CA 

Robert Goyer, University of Western Ontario, National Research Council Board Member/Chair 

Mike Hughes, U.S. EPA, Office of Research and Development, National Health and Ecological 

Effects Research Laboratory, Washington, DC 

Richard Hertzberg, U.S. EPA, Office of Research and Development, National Center for 

Environmental Assessment, Atlanta, GA (Retired) 

Elaina Kenyon, U.S. EPA, Office of Research and Development, National Health and Ecological 

Effects Research Laboratory, Washington, DC 

A-1
Fran Kremer, U.S. EPA, Office of Research and Development, National Risk Management 

Research Laboratory, Washington, DC 

Deirdre Murphy, U.S. EPA, Office of Air and Radiation, Research Triangle Park, NC 

Ecological Effects: 

William Clements, Colorado State University, Ft. Collins, CO 

Heidi Bell, U.S. EPA, Office of Water, Washington, DC 

Peter Chapman, EVS Environment Consultants, N. Vancouver, BC, Canada 

Dale Hoff, U.S. EPA, Region VIII, Denver, CO 

Larry Kapustka, Ecological Planning and Toxicology, Inc., Corvallis, OR 

Michael Newman, The College of William and Mary, Gloucester Point, VA 

Paul Paquin, HydroQual, Inc., Mahwah, NJ 

Cindy Roberts, U.S. EPA, Office of Water, Washington, DC 

Mark Sprenger, U.S. EPA, Office of Solid Waste and Emergency Response, Washington, DC 

Dan Wall, U.S. Department of the Interior, U.S. Fish and Wildlife Service, Washington, DC 

Linda Ziccardi, Exponent, Boulder, CO 

METALS ISSUE PAPER AUTHORS1 

Environmental Chemistry of Metals: 

Donald Langmuir, Hydrochem Systems Corp., Colorado School of Mines, Golden, CO 

Paul Chrostowski, CPF Associates, Inc., Takoma Park, MD 

Rufus Chaney, U.S. Department of Agriculture, Beltsville, MD 

Bernard Vigneault, CANMET Mining and Mineral Sciences Laboratories/Natural Resources 

Canada, Ottawa, ON 

Metal Exposure Assessment: 

Michael Newman, The College of William and Mary/VIMS, Gloucester Point, VA 

Gary Diamond, Syracuse Research Corp., Syracuse, NY 

Charles A. Menzie, Menzie-Cura & Associates, Chelmsford, MA 

Jerome Nriagu, University of Michigan, Ann Arbor, MI 

Ecological Effects of Metals: 

Lawrence Kapustka, Ecological Planning and Toxicology, Inc., Corvallis, OR 

William Clements, Colorado State University, Fort Collins, CO 

Linda Ziccardi, Exponent, Boulder, CO 

Paul Paquin, HydroQual, Inc., Mahwah, NJ 

Mark Sprenger, U.S. EPA, Edison, NJ 

Daniel Wall, U.S. Fish and Wildlife Service, Denver, CO 

Human Health Effects of Metals: 

Robert Goyer, University of Western Ontario, ON 

Mari, Golub, University of California, Davis, CA 

Harlal Choudhury, U.S. EPA, Cincinnati, OH (contributor) 

Michael Hughes, U.S. EPA, Research Triangle Park, NC (contributor) 

Elaina Kenyon, U.S. EPA, Research Triangle Park, NC (contributor) 

Marc Stifelman, U.S. EPA, Seattle, WA (contributor) 

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Bioavailability and Bioaccumulation of Metals: 

John Drexler, University of Colorado, Boulder, CO 

Nicholas Fisher, State University of New York, Stony Brook, NY 

Gerry Henningsen, H&H Scientific Services, Centennial, CO 

Roman Lanno, Ohio State University, Columbus, OH 

Jim McGeer, Natural Resources Canada, Ottawa, ON 

Keith Sappington, U.S. EPA, Office of Research and Development, National Center for 

Environmental Assessment, Washington, DC 

Michael Beringer, U.S. EPA, Region VII, Kansas City, KS (contributor) 

PEER CONSULTATION WORKSHOP PARTICIPANTS 

Breakout Group 1 (Overall Framework): 

Pat Doyle, Environment Canada, Ottawa, ON (breakout group lead) 

Charles A. Menzie, Menzie-Cura & Associates, Chelmsford, MA (Workshop Chairman) 

Bruce K. Hope, Oregon Department of Environmental Quality, Portland, OR 

Breakout Group 2 (Environmental Chemistry): 

Rufus Chaney, U.S. Department of Agriculture, Beltsville, MD (breakout group lead) 

Donald Langmuir, Hydrochem Systems Corp., Colorado School of Mines, Golden, CO 

Mark Barnett, Department of Civil Engineering, Auburn University, AL 

Roman Lanno, Ohio State University, Columbus, OH 

Jim Ryan, U.S. EPA, Office of Research and Development, National Risk Management 

Research Laboratory, Cincinnati, OH 

Breakout Group 3 (Terrestrial Exposure/Effects): 

John Drexler, University of Colorado, Boulder, CO (breakout group lead) 

Lawrence Kapustka, Ecological Planning and Toxicology, Inc., Corvallis, OR 

Bradley Sample, CH2M Hill, Sacramento, CA 

Beverly Hale, University of Guelph, Guelph, ON 

Breakout Group 4 (Aquatic Exposure/Effects): 

Keith Sappington, U.S. EPA, Office of Research and Development, National Center for 

Environmental Assessment, Washington, DC (breakout group lead) 

David Mount, U.S. EPA, Office of Research and Development, National Health and 

Environmental Effects Laboratory, Duluth, MN 

Jim McGeer, Natural Resources Canada, Ottawa, ON 

Steve Klaine, Clemson University, Clemson, SC 

Paul Paquin, HydroQual, Inc., Mahwah, NJ 

Anne Smith-Reiser, Analytical Services Inc., Jessup, MD 

Breakout Group 5 (Human Exposure/Health Effects): 

Steve Devito, U.S. EPA, Office of Environmental Information, Washington, DC (breakout group 

lead) 

Mari Golub, University of California, Davis, CA 

Herman Gibb, Sciences International, Alexandria, VA 

Rosalind A. Schoof, Integral Consulting, Inc., Mercer Island, WA 

Deirdre Murphy, U.S. EPA, Office of Research and Development, Office of Radiation, Office of 

Air Quality Planning and Standards, Research Triangle Park, NC 

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___________________ 

Alexander McBride, U.S. EPA, Office of Solid Waste and Emergency Response, Washington, 

DC 

Margaret MacDonell, U.S. Department of Energy, Argonne National Laboratory, Argonne, IL 

Bob Jones, U.S. EPA, Office of Prevention, Pesticides, and Toxic Substances, Washington, DC 

Richard Troast, U.S. EPA, Office of Solid Waste and Emergency Response, Washington, DC 

1

 Information from the metals issue papers was used in the development of this framework 

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