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# Framework for Metals Risk Assessment

O f fi c e of t h e S c i e n c e A d v i s o r R i s k A s s e s s m e n t F o r u m

### EPA 120/R-07/001 March, 2007



# Framework for Metals Risk Assessment

Office of the Science Ad visor Risk Assessm e nt Forum U.S. Environm ental Protection Agency Washington, DC 20460

### DISCL A IMER
This document has beenreviewed inaccord an ce with U.S. Environm ental Protection Agency policy. Mention of trade nam e s or comm er cial products does not constitute endorsem e nt
or recomm e ndation for use.

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### CONTE N T S
LIST OF TABLES

................................................................................................................. ...... vii LIST OF FI GURES

................................................................................................................ ..... vii LIST OF ABBREVIATIONS AND ACRONYM S

................................................................... 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 Assessm ents

.... ................................................................... 1- 6 1.3.3. Site-Specific Assessm ents

.... ...................................................................... 1- 8 1.4. KEY PRINCIPLES TO CONSI D ER IN METALS RISK ASSESSM ENT

.... ....... 1- 9 1.4.1. Metals are Natu rally Occurring Constituen ts in the Env i ronm ent and Vary in Concentrations Across Geographic Regions

.... ............................ 1- 9 1.4.2. All Environm ental Media have Naturally Occurring Mixtures of Meta ls and Meta ls ar e Of ten Intr oduced into the Environment as Mixtures

..... .............................................................................................. 1- 10 1.4.3. Som e Metals are Essential for Ma intaining Proper Health of Humans, Ani m als, Plants, and Microorganisms

..... ................................................ 1- 10 1.4.4. The Environm ental Chemistry of Metals Strongly Inf l uences Their Fate and Effects on Human and Ecological Receptors

..... ....................... 1- 10 1.4.5. The Toxicokinetics and Toxic odynami cs of Metals Depend on the Metal, the F o rm of the Meta l 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 E C OL OGICAL RISK ASSE SSMENT: PLANNING AND PROBLEM FORMULATION

.... ............................................ 2- 1 2.2. METALS CONCEPTUAL M O DEL

.... .................................................................. 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 Ef fects/Hazard Analy s is

..... ....................................... 2- 14 2.4. RISK CHARACTERIZATION

..... ....................................................................... 2- 16
3. ENVIRONMENTAL CHEMISTR Y, TRANSPORT AND FATE

.... ................................... 3- 1 3.1. INTRODUCTION AND T E RMINOLOGY

.... ....................................................... 3- 1 3.1.1. Hard and Soft Acids and Bases: The Stab ility of Comp lexes

.... ............... 3- 2 3.1.2. Transfor m a tions

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

.... ........................................ 3- 5 3.1.2.2. Abiotic Trans fo rmations

.... .......................................................... 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 o f

p H and Redox Conditions

.... ................................ 3- 9 3.1.4. Ground W a ter and Me tals Mobility

.... ........................................................ 3- 9 3.1.4.1. Applica tion of P a rtition Coe ffi cien ts to Meta l Mobility in Ground Water

..... ....................................................................... 3- 11 3.1.5. Sedim e nt Che mi s try

..... ............................................................................. 3- 13 3.1.5.1. Metal Chemistr y in Sedimen ts ..... .............................................. 3- 14 3.1.5.2. Estimating Me ta l Distr i bution in Sediments

..... ......................... 3- 16 3.1.6. Soil Che mistry

..... ..................................................................................... 3- 17 3.1.6.1. Key Parameters Affec ting M etal Bioa vailability in So ils

..... ..... 3- 18 3.1.6.2. Adsorption Behavior of the Metals of C oncern ..... .................... 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 P l ants
..... ................................................ 3- 20 3.1.7. At m o spheric Behavior/Chemi stry

..... ....................................................... 3- 21 3.2. METAL FATE AND TRANSPORT

..... ............................................................... 3- 23 3.2.1. Aquatic and Terrestrial Tran sport 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 PRINCI PLES

.... ...................................................................................... 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. Nutritiona l Sta te .... ....................................................................... 4- 7 4.2.3.6. Genetic Polymorphisms and Variability

.... .................................. 4- 7 4.2.4. Environm ental Release, T r ansport, and Fate

.... .......................................... 4- 8 4.2.5. Route-Specific D i fferences 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/Toxicodynami cs ..... .......................................................... 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 Toxic ityTests for M etals

..... .... 4- 25 4.3.6. Dose-Response Assessm ent

..... ................................................................. 4- 26 4.4. RISK CHARACTERIZATION

..... ....................................................................... 4- 27 5. AQUATIC ECOLOGICAL RISK ASSESSMENT FOR METALS

.... ................................. 5- 1 5.1. METALS PRINCI PLES

.... ...................................................................................... 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 Toxic odynamics (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/Toxicodynami cs (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 PRINCI PLES

.... ...................................................................................... 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/Toxicodynami cs .... ............................................................. 6- 4 6.2.5.1. Bioavailab ility

.... .......................................................................... 6- 4 6.2.5.2. Bioaccumulatio n .... ...................................................................... 6- 5 6.2.6. Soil Invertebrate Exposure

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

.... ........................................................................................ 6- 7 6.2.8. W ildlife 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. W ildlife Toxicity

..... .................................................................................. 6- 17 6.4. RISK CHARACTERIZATION

..... ....................................................................... 6- 18
REFERENCES

.... ................................................................................................................. ..... R- 1
APPENDIX A: FRAM E W ORK FOR METAL S RISK ASSESSMENT CONTRIBUTORS AND REVIEWERS

.... .................................................................... A- 1

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### LIST OF TABLES Table 3-1. Hard and soft acids (m et al cations) and base s (ligands)

............................................ 3-3 Table 3-2. Examples of or ganom etallic compounds

................................................................... 3-4 Table 3-3. Metals/m etalloids involve d in m e thylation processes

............................................... 3-5 Table 3-4. General trends of enviro nm ental f actors af f

ecting rates of m e thylation/dem e thylation

........................................................................................ 3-6 Table 4-1. Metal essentiality for hum a ns .................................................................................. 4-17 Table 4-2. Summ ary of m a jor differences in kinetic behavior of organic compounds comp ared to m e tals and inorganic m e tal compounds in hum ans ............................ 4-19 Table 6-1. Qualitative bioavailability of m etal c ations in natural so ils to plants and s o il invertebrates

............................................................................................................... 6-8 Table 6-2. Qualitative bioavailability of m etal an ions in natural so ils to plants and s o il invertebrates

............................................................................................................... 6-8 Table 6-3. Metals classified by their known essentiality

.......................................................... 6-13

LIST OF F I GURES Figure 1-1. Categories of m e tals assessm ents under EPA statutory fram ework

........................ 1-5 Figure 2-1. Risk assessm e nt/risk ma nagem e nt process for metals

............................................. 2-2 Figure 2-2. Generic conceptual m o del for m e tals risk assessm e nt

. ............................................ 2-6 Figure 2-3. Conceptual diagram for ev aluating bioavailability processes and bioaccessib i lity for m e tals in so il, s e dim e nt, or aquatic system s ........................... 2-12 Figure 3-1. Approxim a te positions of som e natural environments in term s of Eh and pH

...... 3-10 Figure 3-2. A generalized m ode l fram e w ork for chemical fate and transport in an aquatic system

......................................................................................................... 3-26 Figure 6-1. Generalized repres entation of percent contribution of incidental soil ingestion to oral dose for wildlife atdifferent soil i ngestion rates and bioaccumulation facto r s and a bioavailab ility of 100 percen t

................................ 6-10

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### LIST OF ABBREVIATIONS AND ACRONYMS
1CFOK One-Comp artm ent, First-Order Kinetics m odel ABA Absolute bioavailability aBLM Aquatic Biotic Ligand Model ANZECC Australian and New Zealand Environm ent and Conservation Council ARMCANZ Agriculture and Resources Mana gem e nt Council of Australia and New Zealand ATSDR Agency for Toxic Substances and D i sease Reg i s try AVS Acid-volat ile sulf ide AW QC Am bient W ater Qua lity Crite ria AW QS Am bient W ater Qua lity Standards BAF Bioaccumu lation facto r BCF Bioconcentration factor BF Bioaccess i ble fraction BLM Biotic Ligand Model BSAF Biota/s e diment accumu lation facto r CBR Critical body residue CCA Chrom a ted copper arsenate CEC Cation exch ange capacity CHESS Chemical Equilib ria in Soils and So lutions CMAQ Community Multiscale Air Quality model DEPM Dietary Exp o sure Poten tial Mod el DHHS U.S. Departm e nt of Health and Human Services DL Diffuse Lay e r m odel DOC Dissolved organic carbon DOM Dissolved organic m a tter EC x 

### Effect concentration (for x percent of the study population) ECOFRAM Ecological Committee on FIFRA Risk Assessm ent Methods Eh Redox (reduction-oxidation) potential EqP Equilib rium partition i ng FDA Food and Drug Admi nistration FIAM Free Ion Activity Model FIFRA Federal Inse ctic ide, Fun g icide, and Rodenticid e Act FNB Food and Nutrition Board (NAS) GI Gastroin tes t inal tract HFO Hydrous ferric oxide HLA Hum a n leukocyte antigens HSAB Hard and soft acid and base ICP-MS Inductively coupledplasm a- m ass spectrom etry IEUBK Integrated E xposure Uptake Biokinetic m odel IRIS Integrated R i sk Inform ation System IUR Inhalation R i sk Unit Kd Partition distribution coefficient Kow Octanol-water partition coefficient LC 50 

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

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### LIST OF ABBREVIATIONS AN D ACRONYMS (continued)
MAP Meta ls Action Plan MCL Maximu m c ontaminant level MPPD Multip athway Particle D o sim etry Model NAS National Academy of Sc iences NASQAN National S tream Quality Accounting Network NOM Natural o r ganic m atter NRC National Research Council OM Organic m atter OSWER Office of Solid W aste and Em ergency Response (EPA) PBPD Physiologically based pharm acodynami c PBPK Physiologically based pharm acokinetic PBT Persis tent b i oaccumu lative toxic PBTK Physiologically based toxicokinetic QSAR Quantitative struc ture - ac tivity re latio n ship RAGS Risk Assessm e nt Guida nce for Superfund RBA Relativ e bio availability RDA Recommended dietary allowance RDDR Regional Deposited Dose Ratio RfC Reference concentration RfD Reference dose RTC Report to C ongress SAB Science Adv i sory Board SEM Simu ltaneously extrac ted m etals SHEDS Stochastic H u m a n Exposure and Dose Simulation m odel SPC Science Policy Council (EPA) SRWG Science and Research Working Group (SRW G) of the Non-Ferrous Metals Consultative Forum on Sustainable Developm ent STORET Storage and Retrieval data system TCLP Toxicity Ch aracteristic Leaching Procedure TMDL Total Max imum Daily Load TRIM Total Risk Integrated Methodology TRV Toxicity ref erence v alue TSCA Toxic Substances Control Act U.S. EPA U.S. Environm ental Protection Agency W E R W a ter-effect ratio WHAM W i nder m ere Humi c Aqueous Model WHO World Health Organization WHO/IPCS World Health Or ganization/International Progr amme on Che mical Safety

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### PREF ACE Many U.S. Environm ental Protection Agency (EPA or the Agency) program s are designed to develop guidelines on ho w to regulat e m etals. I n this process, decisions can range from setting environm ental releas e standards, to establishing pr otective levels in different environm ental m e dia, to setting priorities for pr ogramm atic or voluntary e fforts. A fundam e ntal input to the decision-m aking pro cess for m ost EPA program s is an assessm e nt of the potential risks to human health an d the enviro nm ent. EPA’s Science Policy Council (SPC) recognizes that m e tals present unique risk
assessm ent issues. The SPC tasked an Agency workgroup, under the auspices of EPA’s Risk Assessm ent Forum , the challeng e of devising a Metals Action Plan (M AP) to estab lish a p r oces s for ensuring the consistent applicatio n of scientif ic princ i ple s to m etals ris k assessm ent. The MAP included brief descriptions of the Agency’s curren tactivities on m e tals, identif i ed critic al scientific issues that need addressing, and recomm ended the developm ent of a m etals risk
assessm ent fram ework. The MAP stated that th e Fram ework should offer general guidance to EPA program s for considering the various properties of m e ta ls, such as environm ental chemistry, bioavailability, and bioaccumulation. Because of the scien tific co m p lexity of m etals-sp ecific risk as sessm ent, the Agency recognized the need to include stakeholders and the public in the Fram ework developm ent process and to involve experts throughout the A gency. A stepwise process was initiated, beginning w ith the creation of MAP and continui ng with fram e w ork developm ent and review. Workshops and peer-review activities were conduc ted at regular intervals during developm ent of the Fram ework to ensu re that curren t, accu rate s cience supp orts th e prog ram applications. To gain additional inform ation, the Agency contr acted for the developm ent of issue papers on im portant topics in m etals assessm ent. These activities, along with input f r om other federal agencie s an d review by EPA’s Science A dvisory Board (SAB), provided a ddition a l improvem ents to th e Fra m ework draf ting proces s. Addition al details on these a ctiv ities are provided below. MAP Stakeholder Input. In February 2002, MAP convened a m eeting to gather stakeholder input to help EPA fo r mulate the plan for developing the Fram ework for Metals Risk Assem e nt. EPA solicited input on organization and content and received comments that were adopted into the docum ent with the greatest practical extent. The m e eting report and comm ents are availab l e on EPA’s Web site at http://cfpub2.epa.gov/ncea/raf/r ecordisplay.cfm ?deid=51737 and http://cfpub2.epa.gov/ncea/raf/ recordisplay.cfm ?deid=51736. Science Advisory Board Review . I n Septem ber 2002, EPA’s SAB reviewed the MAP and provided comments. By this point, MAP ha d som e of the panel’s recomm endations are

x

### summarized below; the f ull text of al l recomm endations is av ailab l e at http://www.epa.gov/sab/pdf/ecl03001.pdf.

• The panel agreed that inorganic m e tals s hould be assessed differently from organic pollutants in a num ber of contexts. Meta ls are elem ents and, although they do not degrade, they have comp lex environm ental chemistry. Moreover, som e m e tals are essential for living organism s, and metals occur n atura lly in th e environm ent.

• The panel agreed th at chemical speciation, bioavailab i lity, b i oaccumu lation, and toxicity are key issues inassess ing th e hazard s of m e tals, with qualificatio ns.

• The panel recommended consideration of stab ility and environm enta l residence tim es, as well as o v erall enviro nm ental che mistry, to de termine temporal ch arac teri stics of m etal hazards.

• The panel recommended greater emphasis on th e com b ined effects of m etals, including nutritiona l and toxicological considerations. Issue Paper Topics and Science Questions. As part of the development of the Framework, issue papers were deve loped 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 re search needs. The papers, available at http://cfpub.epa.gov/ncea/raf/recordisplay.cfm ? deid=86119, address the following topics and primary questions:

• Environmental chemis try. How can environm ental ch emistr y be better incorporated into ass essments f o r inor ganic m etals ?

• Bioavailability and bioaccumulation of metals. Wh at m ethods or tools can be used now to ref l e ct m etal bio availability ? Wh at scien tif ica lly bas ed approach es can be used to determine m etal bioaccumu lation ?

• Metal exposure assess ment. What are the relevant expos ure pathways for inorganic m e tals to hum ans and ecological endpoints?

• Human health effec t s. What populations are m o st su sceptible to effects from inorganic metals? How should toxicity test s be conducted and inte rpreted, including issues of essential elem e nts and dietary salts, am ong others?

• Ecological effects. Wh at ecolog i cal system characteris tics p r om ote increased toxic ity f ro mm e tals ? Peer Consultation Workshop. A draft fram e w ork was comp leted in July 2004, and a peer consultation workshop was held from July 27-Ju ly 28 to seek input from expert scientists in the m e tals risk assessm ent field. W o rkshop participants were from academia; industry, state, federal, and Canadian agencies; and various EP A program offices. Stakeholder comments were xi

### also received for consideration. Based on co mments received at the workshop, the Agency contracted with a few workshop participants to explore several issues and insufficiencies identified in the hum an health and en vironm en tal chemistry discussions. The Fram ework draft was revised and m ade available for inter-Agency review. Inter-Agen cy Review . Based on comments received, the fram e w ork was revised. SAB Review and Public Comment. The SAB review was conducted in February 2005, and the final draft of the Framework for Me tals 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/m etals_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 asse ssment issues rather than tools.

xii

### STEE RING COMMITTE E AND CONTRIB U TORS 1


### CO-CHAIRS Anne Fairbrother, U.S. EPA, Office of Res earchand Develo pm ent, National Health and

Environm ental Effects L aboratory, We stern Ecology Division, Corvallis, OR Randall W e ntsel, U.S. EPA, Office of Re search and Developm ent, W a shi ngton, DC STEERING COMMITTEE Stephen DeVito, U.S. EPA, Office of E nvironm e n tal Inform a tion, W a shington, DC Alexander McBride, U.S. EPA, Office of So lid Waste and Em ergency Response, W a shington, DC (Retired ) David Mount, U.S. EPA, Office of Researchand Developm ent, National Health and Environm ental Effects L aboratory, Duluth, MN Keith Sappington, U.S. EPA, Office of Resear ch and Developm ent, National Center for Environm ental Assessment, W a shington, DC FORUM STAFF Gary Bangs, U.S. EPA, Office of the Science Advisor, Risk Assessm ent Forum , Was hington, DC Marilyn Brower, U.S. EPA, Office of the Science Advisor, Risk Assessm ent Forum , Washington, DC Elizabeth Lee Hofm ann, U.S. EPA, Office of th e Science Advisor, Risk Assessm ent Forum , Washington, DC Pa m e la Noyes, U.S. EPA, Office of the Scien ce Advisor, Risk Assessm e nt Forum , Washington, DC Seem a Schappelle, U.S. EPA, Office of the Scien ce Advisor, Risk Assessm e nt Forum, Washington, DC W illiam Wood, U.S. EPA, Office of the Scien ce Advisor, Ris k Assessm e n t Forum , Washington, DC CONTRACTOR SUPP ORT FOR FINAL DOCUMENT Patrick J. Downey, FTN Associates, Inc., Little R ock, AR Lynne Haber, Toxicology Excellence for Risk Assessm ent, Cincinnati, OH Ingeborg Harding-Barlow, Palo Alto, CA Charles A. Menzie, Menzie-Cura & Associates Inc., Chelm s ford, MA Kent Thornton, FTN As sociates, Inc., Little Rock, AR

Meta ls issu e paper au tho r s and peer consultation workshop partic ipants are listed in Appendix A.

1 Con tribu tors an d rev i ewers are listedin App e nd ix A. 

### xiii

### EXE CUTI V E SUMM ARY

The Fram ework for Metals Risk Assessm ent is a science-based docum ent that addresses the special attributes and beha viors of m e tals and m e tal comp ounds to be considered when assessing their hum an health and ecological risks. The docum ent de scribes basic principles to be considered in assessing risks posed by m e tals an dis intended to f o ster con sisten c y in h o w these principles are applied across the Agency’s program s and regions when conducting these
assessm ents. Although the audience for the Framework is prim arily intended to be Agency risk
assesso rs, it will a l so co mmunicate princ i ples and recomm endations f o r m etals risk a ssessm ent to the s t akeh olders and the public. T h is guidan c e will be us edin conjunc tion with guid a nce developed by the program s and regions for use in site-specific risk assessm e nts, criteria derivation, ranking or categorization and other similar Agency activities related to m e tals. The purpose of this docum ent is to present key guiding principles based on the unique attributes of m e tals (as differentiated from organic and organom e tallic compounds) and to describ e ho w these m etals-sp ecific attributes an d pr inciples m a y then be applied in the context of existing EPA risk assessm e nt guidance and prac tices. Wh ile organic compounds, for example, undergo bio accumu latio n, there are unique prop erties, issues, and pro cesses within th ese principles th at assessors need to consider when evaluating meta l compounds. Furthermore, the latest scientific data on b i oaccumulation do not currently sup port th e use of bioconcentration factors and bioaccumu lation facto r s when applied as generic thresho l d criteria for the hazard potential of m e tals. W h ile the sc ience su rrou nding the m etals risk assessm ent principles continues to be studied intensively and evolving rapidly, som e areas s t ill lack sufficient infor m ation for a quantitative assessm ent to be carried out. Thus , specific approaches m a y becom e outdated or
m a y otherwise require m odification to reflect th e best availab l e scien ce and others m a y be addressed o n ly qualitatively until ad dition al inf o rm ation becom es available. Regard less, th e following principles are more generalized, fundam e ntal properties of m e t a ls and should be addressed and incorporated into all i norganic m e tals risk assessm ents:

• Meta ls ar e n atura lly oc c u rring con stitue nts in the environm ent and vary in concentrations across geographic regions.

• All environm ental m e dia have naturally occu rring mi xtures of m e tals, an d m etals ar e of

ten introduced into th e environm ent as mi xtures.

• Som e m etals are es sential f o r m aintaining pr oper health of hum a ns, ani m a l s, plants, and mi croorganism s.

• Metals, unlike organic chemicals, are neith er created nor destroyed by biological or chemical processes; although, these processe s can transform m etals from one species
xiv

### to another (v alence states ) and can co nve rt them between inorganic and organic for m s.

• The absorption, distribution, transform a tion, and excretion of a m e tal within an organism , depends on the m e tal, the for m of the m e tal or m e t a l compound, and the organism ’s ability to reg u late and/or store the m e tal.
This Fram ework document is not a prescripti ve guide on how any particular type of
assessm ent should be conducted within an EPA program office. Rather, it is intended to outline key m e tal principles and how they should be cons idered in existing hum an health and ecological risk as sessm ent practices to foster co nsiste ncy across EPA program s and regions. Chapter 1 identifies the purpose, audience, and scope of the Fram ework; includes a discussion of typical programm a tic or regulatory contexts under whic h the Agency mi ght undertake a m e tals risk
assessm ent; and provides an overview of the m e tals principles. In particular, the context for risk
assessm ent is an im portant factor in determining th e type of analysis appr opriate for a particular situation. T o provide a context for discussion of the fram ework principles for m e tals, EPA has defined three general categories of assessm ents: national ranking and categorization, nationallevel assessm e nts, and site-specific assessm ents. Each type of assessm e nt can vary in level of detail f r om sim p le, scree n ing analy sis to com p lex, def i nitiv e a ssessm ents. Approaches and
m ethodologies m ay be appropriate f o r som e or all the risk assessm ent types and m ay be m o re or less app licable to s creen ing or com p lex assessm ents. Here an d elsewhere, the Fram ework acknowledges that data m a y not be available to imple m ent all the step s in a m e tals risk
assessm ent (e.g., lack of inform ation about m e ta l speciation in som e environm ental media), requiring use of assumpt i ons and a discussion of how such uncertainty influences the risk outcom e . Chapter 2 reviews the hum a n health and ecol ogical risk assessm ent paradigm and applies the m etals p r inciples to each phase o f the risk as s essm ent process using assessm ent questions. The chapter presents a metals concep tual m odel showing the interrelations hip between the m etals
or m e tal compounds of interest and the assessment process. Chapter 3 describes how environm ental chemistry issues affect the assess m e nt of metals. Chapter 4 describes how the
m e tals principles should be considered when conducting hum an health risk assessm ents, and Chapters 5 and 6 describe how the metals princi ples should be considered in the context of aquatic and terres trial ris k assessm ents, resp ectiv ely. This document discusses scientif ic issues and scientific appr oaches. It is intended to foster the consistent application of methods and data to m e tals ri sk assessm e nt in consideration of

the uniqu e propertie s of m etals. The Agency will be an alyzing the science policy implications and developing appropriate policy approaches th at are protective of hum a n health and the environm ent.

xv

### The Fram ework is th e result of contributions from a variety of individuals inside and outside the Agency. Their com b ined expertis e and enthusiasm have impr oved the technical quality of the docum ent and its applicability for various risk assessm ent activ ities.

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### 1. INTRO DUCTION Inorganic metals and m e tal compounds have unique characteristics that should be considered when assessing their risks. Som e of t hese characteris tics typically are not considered when assessing the risks of organic substances. For example, m e tals are neither created nor destroyed by biological or chemi cal processes; they are trans f or m e d from one chemical form to another. Native (zero valence) form s of m ost m e tals and som e m e tal compounds are not readily soluble, and as a result, toxici ty tests based on soluble salts m a y overestim ate the bioavailability and toxicity of these substances. Som e metals (e.g., copper [Cu], seleni um [Se ] , and zinc[Zn]) are nutritionally essential elem ents at low levels but toxic at hi gher levels, and others (e.g., lead [Pb], arsenic [As], and mercury [Hg]) have no known biological function s. Because m e tals are naturally occurring, m a ny organism s have evol ved m echanism s to regulate accumu lations, especially a ccumu lations of essentia l m etals. Th is m etals ris k assessm ent Fram ework identif i es m e tals principles that are funda m e ntal truths (or properties) of m e tals. The m e tals principles should be addressed and incorporated into inorganic m e tals risk assessm ents.

Metals Principles

• Metals a r e na turally occu rri ng co nstitue n t s in the environment an d vary in con c e n trations across ge og raphi c regi on s.

• All environme n tal media ha ve naturally o c curri ng mixture s of metal s , and metal s are often introdu ce d into the environ ment as mixture s
.

• Some metals are e s sential for maintaini n g prop er h eal th of humans, animals, pla n ts, and microorgani sms.

• The environm ental ch emist ry of metals strongl y influe nce s their fat e and effect s on huma n and e c olo g ical recepto r s.

• The toxico kin e tics a nd toxicodyn a mics o fmetals depe nd on the met a l, the form of the metal or metal com poun d, and th e orga nism’s ability to regu late and/or store the metal.



### Because the m ajority of com pounds asse ssed by the United S t ates Enviro nm ental Protection A gency (EPA or the Agen cy) are organic substances, the various guidance docum ents provided for risk as sess m e nts of either hum an h ealth o r eco logical recep tors lack specificity on how to acco unt for these and other m e tal attribu t es. This docu m ent attem p ts to fill th is gap in current guidance.
1.1. PURP OSE AND AUDI ENCE The prim ary purpose of the Fram ework for Metals Risk Assessm e nt is to identify key principles that should be addressed in any inorganic m e tals analysis and to provide EP A program 1-1

### offices and regions with guidance on how to cons ider these principles in EPA risk assessment practices. Although the primary audience will beAgency risk assessors, the Framework will also communicate these principles to stakeholde rs 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 meta ls. The papers are available on EPA’s Web site at http://cfpub.epa.gov/ncea/raf/recordisplay.cfm ? d eid=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 develope d to supplement previous guidance for use in site-specific risk assessment s; criteria derivation, ranking, or categorization; and other similar Agency activities related to metals. The Fr amework 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 freque ntly encountered when conducting a metals-specific risk assessment. This documen t 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 asse ssments of risks from chemicals in general. Information on general risk assessment topi cs is available on EPA’s Web site at http://cfpub.epa.gov/ncea/ and http ://cfpub.epa.gov/ncea/raf/index.cfm. The Fram ework is intended to be used fo r guidance only. It does not establish any substantive “rules ” under the Admi nistrative Pr oc edure Act o r any other law and will h a ve no binding effect on EPA or any regulated entity. Rather, it represents a nonbinding statem ent of policy. EP A believes that the Fr amework pr ovides a sound, up-to-date presentation of principles; provides guidance on how to consid er these principles in assessing the risk posed by
m e tals; and enhances application of the best av ailable science in Agency risk assessm ents. However, EPA m a y conduct m e tals risk assessm en ts using approaches th at differ from those described in the Fram ework for m a ny reasons , including, but not limite d to, new inform ation, new scientific understandings, and new scie nce policy judgments. Wh ile the science surrounding m e tals risk assessm ent continues to be studied intensively and is evolving rapidly,
som e areas still la ck suf f i cient inf o rmation f o r a q u antitative assessm ent. Thus, specific approaches m a y becom e outdated or m a y otherwise require m odi fication to reflect the best available science and others m a y be addressed only qualitatively until additional information becom es availab l e. Ap plication of this Fram ework in f u ture m etals ri sk assessm ents will be based on EPA decisions that its app r oaches ar e s u itab l e and appropriate. These judg m ents will be tested and examined through peer review, and any risk analysis will be modified as deem ed appropriate.

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### 1.2. METALS FRAMEWORK SCOPE The Agency regulates m e tals and their ino r ganic and organo m e tallic co mpounds (compounds exhibiting properties of both organic and m e tal compounds) because they have the potential to harm hum a n health and the environm ent. The Agency’s SAB has stressed Metals and Metalloids of Primar y Interest 

### the importan ce of enviro nm ental che mistry Aluminum Manga ne se 

### and its relevance to the assessm ent of both Antimony Mercury (i norgani c) 

### inorganic and organom etallic comp ounds. Arse nic Barium Molybden um Nickel 

### However, the complexities of addressing all Beryllium Selenium 

### types of m e t a l compounds within a single Boron Cadmi u m Silver Strontium 

### docum ent would result in a fram ework that Chromium Cobalt Tin Thallium 

### would be difficult to follow or to apply in Cop per Vanadiu m 

### specific c ase s. Because organom etallic Iron Lead Zinc



### compounds exhibit properties common to both organic substances and m e tal
compounds, the properties of both the organi c m o ieties of these compounds and their
components would need to be addressed. EPA ha s already developed fram e w orks and associated guidance docum e nts for assessing properties of organic comp ounds. Therefore, this docum e nt addresses only the assessm e nt issues asso ciated w ith inorganic m e tal comp ounds. The Fram ework does discuss natural transfor m a tion pa thways that for m organom e tallic compounds and refers the reader to appropriate A gency docum entation or research efforts related to relevant risk assessm ent issu es. In this Fr amework, the term “m etals” refers to in organic m etals and m etalloids thatmay pose a tox i c hazard and are currently of prim ary in teres tto EP A. However, the principles and approaches set forth in the Fram ework are applicable to all m eta ls. In som e instances, m e tal-bym etal consid erations are included, either as ex amples or as ways to highlight particular exceptions. Furtherm ore, in som e cases, this doc um ent m a y discuss particular tools or m e thods that exp and on a particular prin cip l e and its cons ideration in the contex t o f EPA hazard and risk
assessm ent. 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 m e thod, although proper citations to the open scientific literature are included. Nor does the Fram ework provide an exhaustive summary of all the tools and m e thods availabl e to risk assessors, as this type of analysis is beyond the scope of this docum ent.

1.3. METALS ASSESSMENT CO NTEXT The context for the risk assessm ent is a m ajor factor in determining the type of analysis appropriate for any particular situation. The Agency conducts a variety of assessm e nts, from

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### site-specific risk assessm ents to national criteria setting and ranking. To provide a context for discussion of the Fram e w ork principles for m e tals, th is doc um ent has defined three general categories of m etals assessm ents: national ranking and categorization, national-level
assessm ents, and site-specific assessm e nts. (S ee Figure 1-1 identifying th e three categories of
assessm ent in the contex t of the Agency’s statu t o ry authority.) As shown in the figure, nationalscale and site-specific as sessm ents can vary in le vel of detail—from si mple screening analyses to
comp lex definitive assessm e nts. For example, in conducting a national-le vel screening analysis, EPA mi ght undertake a screening-le vel review of a pesticide or new chemical under the Federal Insecticide, Fungicide, and R odenticide Act (FIF RA) or the Toxic Substances Control Act (TSCA), and this m a y lead to a m ore definitive assessm ent if the screen cannot rule out a threat to health or the environm ent. Site-specific sc reening-level assessm ents mi ght be sufficient to support an environm ental im pact statem ent, and a m o re com p l ex analysis mi ght be necessary as a part of a S uperfund cleanup action. All three categories of assessm ents share common elem ents and rely on accurate inform ation a nd knowledge about how metals behave in the environm ent and when they com e into contact w ith hum ans or other organism s of concern. Metals have unique environm ental and toxico logical properties that m a y confound such
assessm ents if they are not given consideration. Each of the three general assessm ent categories is discussed in m o re detail below.

1.3.1. National Rankin g and Categoriz ation EPA m a y rank or catego r ize som e chemicals ba sed on their potential to cause risk to hum a n health or the environm ent. Although there continue to be gaps in data to understanding the chemistry, environm ental behavior, toxicity, and exposure potential f o r m a ny chemicals, EPA is taske dwith prote cting and mitigating exp o sures and h armf u l ef f

ects assoc i at edwith exposure to these chemi cals. The A g ency of ten is in a position, despite imperfect and incomp lete databases, where m e thods and tools n eed to be developed to identify, characterize, and in som e cases, rank and catego r ize chemicals. W ith m o re than 80,000 chemicals currently list ed on the TSCA inventory that can legally be used in comm erce within the U.S. (not includi ng pesticides or chemicals that are created as byproducts during industrial proces ses), the Agency needs a way to prioritize substances for review or action. Many of the st atutes admi nistered by EPA provi de specific lists of chemicals that require consideration, but of ten those lists are based on infor m ation and analyses previously developed by EPA. In addition, the statutes ge nerally provide for adding or deleting chemicals f r om the initial lis t on th e basis of th eir po tential threat to hum a n health or ecological receptors. Consequently, a need exists for m e thods that rapidly screen chemicals for placem e nt on lists or that p r ioritize potentially hazardous substances.

1-4Categories 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.

Som e of the ranking and categorization m e thods used by EPA involve identifying certain attributes of chemicals th at can then b e used as indicators of potential hum an health or ecological risk. Example attributes include hum an and eco logical toxicity, producti on volum e, quantities rele ased to the environm ent, pers is te nce in th e en vironm ent, mobility in the environm ent as indicated by volatility or solubi lity, and potential to bioaccumulate in the f ood chain. Other
m e thods, which m a y be less quantitative, rely m o re on a com bination of expert judgment, stakeholder input, and availability of infor m ation to determine th e priority or categorization of chemicals for decision making or other action. E x amples of program s where EPA identifies or catego r izes chemicals for prio rity action based onhum a n health or ecol ogical concerns include the f o llowin g :

• Selecting ch emicals for the Agency’s Toxicity Ch aracteristic regulation (4 0 CFR 261.24) that defines hazardous wastes

• Establishing reporting thresholds for sp ills of hazardous m a terials under S uperfund

• Setting p r io rities f o r rev i sions to the Am bient W ater Qua lity Crite ria (AW Q C)
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### • Listing chemicals under the Toxics Release Inventory

• Determining priorities for deve loping drinking water standards

• Setting priorities for hazardous air polluta nt data collection and assessm ent, and

• Setting priorities for reviewi ng existing chemicals under TSCA.
This lis t of needs f o r ranking or categorizing chemicals is not comprehensive but is illus tr ativ e o f the activ ities that EPA conducts in this regard. In addition, the Agency m a y set national standards and guidelines for specific chemicals, including
m e tals, as described in the next section.
1.3.2. National-Level Assessments National-lev el asse ssm ents
m ay be perform ed when the Agency is setting m e dia standards or guidelines for chemicals (e.g., Maximu m Contaminant Levels [MCLs], Na tional Am bi ent Air Quality Stan dards, AW QC, Superfund soil-screening levels) or when the Agency is us in g risk
assessm ents to establish controls for environm ental releases from industry or other sources (e.g., hazardous w aste listings under the Resource Conservation and Recovery Act, residual risk determinations under the Clean Air Act, and pestic ide registrations). These assessments can vary in level of detail from simple, screening an alysis to complex, data -intensive definitive assessments. EPA also is charged with establis hing controls on environmental releases based on the best available treatment technologies (e.g., maximum achievable control technology for air Ambient Water Qualit y Criteria EPA’s Offic e of Water is c h arged with developing Ambient Wat e r Quality Criteria (A WQ C) to supp ort the Clean Wate r Act go als of prote c ti ng and mai n taining the p h y sical, chemi c al, an d biologi cal in tegrity of U.S. waters. Examples of chemi c al -specifi c crite r ia in clud e those desi gne d to protect human h ealth , aquatic life, and wil d life. Although AWQC a r e typically deriv ed at a nation a l level, there is a long hi story behin d the de velopment of method s to accommo date sitespecific differences in met a ls bioavailability. For example, sin c e the 19 8 0 s, aqu atic life crite r ia for several cationi c metals have been exp r e s sed as a fun c ti on of water h a rdn e ss to addre s s the combi ned e ffect of certai n cation s (p ri ncip ally cal c ium an d magne siu m ) on toxicity. Reco gni zing th at water hard ne ss a dj ustment s did not accou n t for othe r impo rtant ions a nd liga nds that can alter metal s b ioavailability and toxic ity, EPA developed the water effec tratio (WE R) pro c ed ure as an empi rical approa ch for makin g site -specifi c bioavailabilityadjustme n ts to criteri a (U. S . EPA, 1994a). This approa ch relies on comp aring toxicity measure m ent s made in site wate r with those made in lab o rat o ry wate r to derive a WER. The WER is then u s ed to adju s t the nation a l criterion to reflect site-specific bi oav a ilability. More recently, the Office of Wat er ha s bee n d eveloping a mech ani stic-based approa ch for addressin g m e tals bio a vailability using t he Biotic Ligan d Model (BL M ) (Di To ro et al. , 2001; Santo r e et al., 2001; U.S. EPA, 2000b). This model , discussed in Cha pter 5, predict s acute toxici ty to aquatic org ani sm s on the basi s of physi cal an d ch em ical facto r s af fecting spe c ia tion, compl e xation , and comp etition of metals for intera ctio n at the biotic ligand (i .e., the gill in t he case of fish). The BLM has been mo st extensively dev elope d for co pper a nd is b eing inco rpo r ate d dire ctly into the nation a l coppe r aqu atic life crite r ion. The BLM is also being d e velo ped for u s e with other metals, incl ud ing silver. Co nce p tually, the BLM has a ppeal becau se met a ls criteri a co uld be imple m ented to accou n t for predi cted peri ods of enhanced bi oavailability at a site that may not be capture d by pu rely empiri cal methods, su ch a s the WER.



### 1-6

### emission and best available treatm e nt techno logy for surface water discharges and for hazardous wastes). However, even though the standard s are based on technologi cal achievability, the Agency typically perform s risk assessm ents in support of these regulations to help Ha zardo us Was t e Lis t ing Determination Und e r the Re sou r ce Co nservation an d 

### inform managem ent decisions and for use in Recovery Act , EPA is required to make formal 

### cost/b enefit analyses. deci s io ns o n wheth er to de sign ate ce rtai n spe c ific indu st rywa st e st rea m s a s hazard ou s
. For wa ste 

### Differing environm ental conditions strea m s that are liste d as hazard ou s , the 

### across the country affect the gene rato rs a nd han dlers o fthose wa ste s mu st comply with a comp reh e n s i v e set of man ageme n t 

### biogeochemi stry of m e tals, m aking it and treatm e n t standa rds. In determi ning whethe r 

### dif f i cult to s et sing le-v alue nation al c r iteria to list a waste as ha zardou s, the Agen cy evaluates the ways in which that wast e is cu rrently being 

### that repre se n t the s am e risk level ac ross the manag ed o rcould pla u siblybe manag ed by the 

### whole country (national standards that apply gene rato rs a nd han dlers o fthe waste. The Agen cy also a s se sse s the physi cal and ch emica l 

### at the point of exposure, such as MC Ls, are comp ositio n of the waste. Based on th e wa ste characteri stics and manag ement practi ces, EPA 

### less affected by these factors). To conduct then co ndu ct s an an alysi s to determin e wheth er 

### such asses s m ents, the Agency commonly potentially ha rmful co nstitu ents in the waste might be rele ased a nd tran spo r te d to human o r 

### undertakes several approaches. One is to ecol ogi cal re ceptors. In co ndu cting the s e 

### define one or m ore exposure scenarios and analyses, the Agency eval uates the p o tential for con s tituent s in the wa ste material to be relea s ed to 

### to conduct a relatively detailed analysis. air, su rface water, soil, and groun d wate r. It then 

### The dif f i culty in this app r oach is in s elec ting model s the fate and tran sp ort of those constitue n ts to potential re cepto r s. 

### the approp riate scen ario; typica lly, th e Agency tries to ensure th at the scenario is sufficiently conservative to be protective of the population at highest risk(such as populations exposed above the 90th percentile) without be ing so conservative th at the standards are protective of hypothetical individuals whose calculate d risks are above the re al risk distribution. In selecting the appropriat e scenario, the Agency needs to con sid er all the factors th at m ay affect potential risk, including environm ental factors affecting the fate, transport, exposure potential, and toxicity of the chemi cals released

. Another common approach for a national assessm e nt or criteria derivation is to conduct a probabilis tic analys is (su ch as a Mon t e Carlo analys is), wher ein the vari ability of the k e y f actors is desc ribed by param eter distribu tio ns used as in puts to th e p r obability an alysis p r oce dure. The result is an integrated distribu tion of potential risk levels. Th e difficulties related to conducting this kind of analys is are in develop i n g appr opriate distribu tio ns for each param e ter and in ensuring that adequate attention is paid to potential correlatio ns am ong ke y param e ters. These correlations often are m o re complex and difficultto describe for m e tals than for organic
compounds.

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### 1.3.3. Site-Specific Assessments Site-specific assessm ents are conducted to inform a decision concerning a particular location and m a y also support som e national regulato ry decisions. They can also vary in detail f r om screening-leve l to c o mp lex, def i nitiv e-level analyses.

• Determining appropriate soil clea nup levels at a S uperfund site • Establishing water discharge permit conditions to m eet AWQS, and • Determining the need for emission standard s for sources of h azardous air pollutants.

Establishing Water Disc h a rge Permit Conditio ns The Cle an Water Act esta blish e s a two-tier pro c e s s for setting wat e r disch a rge permit conditi ons. First, all disch a rge r s must meet the tech nology-ba s ed effluent guid e line s limitations requi rem ents. Secon d , if those limitation s are not ad eq uate to a llow the receiving strea m to ach i eve its desi g nated water q uality stand ard s (WQS), then mo re stri nge nt limits are d e ve loped to en su re that those stand ard s a r e met. WQS are esta blishe d by the st ates and con s i s t of a desig nat ed use for the water body and a set of criteria f o r individual chemical s thatallow that use to be achi eved. EPA ha s publi s hed n a t i onal water q uality criteri a values for the states to use as guid a n c e in setting thei r stand ard s
. Once the sta ndards that in clud e the crit eria have b e e n establi s h e d and it has be en determine d that the effluent guideline s alone will not be suffici ent to allow tho s e crite r ia to be met, the state prep ares a wastelo ad allo cation for all the discha rge r s to that strea m segm ent, inclu d ing, whe r e approp riate, the non point source di scha rges. The wa steloa d allo ca tion gene rallycon s ist s of modelin g the potential i m pact on the strea m from each disch a rge of the che m ical s of con c ern and the n setting the allowable discha rge s to ensu r e that the crite r ia for the chemi c al s are met. The modelin g proce s s can b e quite compl e x, potentially taking into account the intera ctio ns of the amb i ent strea m condition s with the ch e m ical s in the discha rge, in cludi ng dilu tio n , chemi c al transfo rmatio n s , degradatio n, settling, resuspen sio n , and other p r oce s se s. For metals, str e a m cha r a c teri stics such as p H , orga nic co ntent, su spe nde d solids level s , a nd nume r o us other fa ctors can signifi can t ly affect how the metal will behave and affect aq uatic life in the strea m se g ment. Theref ore, it is impo rtant to unde rstand the s e pro c e s ses in con du c ting th e wa steloa d allocation.



### An accurate site-specific assessm ent for aninorg anic m etal requires know ledge of the for m (or form s) of the metal as it enters the environm ent, the environm ental conditions affecting the m e tal (clim a tological conditions, soil geochemi st ry, water and sedim e nt chemistry, etc.), the existence of plants and/o r anim als in which th e metal mi ght bioaccumu late as well as the uptake factors for whatever form (s) the m e t a l m a y be in, plausible pathways and routes of exposures to the hum an or ecolog i ca l recep tors, an d the effect the m etal will have on target organ i s m s in whatever form in which it reaches that organi sm a nd its target organ/system. Although m a ny of these sam e principles als o affect the risk poten tial of organic chemicals, models for predicting fate, transport, and toxic properties ar e genera lly bette r defined for organic chemicals than for
m etals.

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### 1.4. KEY PRINCIPLES TO CONSIDERI N METALS RISK ASSESSMENT One of the purposes of this Fram ework is to present key principles thatdif f

erentiate inorganic metal com pounds from other chemicals. These key principles, defined in subsequent subsections, warrant careful consideration when assessing the risks to human health an d the environm ent associated with exposures to m e ta ls or m e tal compounds and should be addressed and incorporated into m e tals risk assessm ents to the extent pract icable. F o r exam ple, it is known that certain m etal com p ounds bioaccumulate in hu m a n tissues and that this bioaccumulation can be related to the m etals’ toxic ity (SA B , 2006). Contributors to the Metals Action Plan (MAP),
m e mbers of the SAB, and extern al stakeholders, along with vari ous contributors to and authors of this Framework, have discusse d th ese m etals p r inc i ples f o r consider atio n in the a sse ssm ent of
m e tals. They are visible throughout this docum ent. In Chapter 2, they are broadly discussed in the context of the Agency’s risk ass e ssm ent proce ss. In Chapters 4 through 6, they are discussed in the context of hum a n health, a quatic, or terrestrial risk asse ssm ent processes. The rem ainder of this chapter discu sses these principles in m ore detail, focu sing on the unique properties of
m e tal comp ounds and why these principles are important for m e tals risk assessm ents.

1.4.1. Metals are Naturally Occurring Constituents in the Environment and Vary in Concentrations Across Geographic Regions Implications f o r risk ass e ssm ent inclu d e the f o llo wing:

• Hum a ns, other anim als, and plants have e volved in the presence of m e tals and are adapted to various levels of m etals. Many anim als a nd plants exhibit geographic distributions that reflect vari ab le r eq u irem ents f o r and/or tole rance to certain m etals. These reg i o nal differences in requ irem ents and tolerances should be kept in mi nd when conducting toxicity tests, evaluating ri sks, and extrapolating across regions that dif f

er natura lly in m etals levels.

• As a result of industrialization, current leve ls of metals m a y be elevated relative to levels occurring naturally. Depending on th e purpose of the risk assessm ent, care should be taken to understand and disti nguish among naturally occurring levels, current background levels (i.e., natural and anthropogenic sources), and contributions to curr ent le vels f r om speci fic activities of concern.

• Because the diets of humans and other an im als are divers e, th ere m ay be wide variability in the dietary intake of som e metals (e.g., in seaf ood), resulting in both temporal variability (e.g., spikes after a seafood meal or with life stage) and geographic or cultural variability.

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### 1.4.2. All Environmental Media have Naturall y Occurrin g Mixtures of Metals and Metals are Often I n troduced into the Environment as Mixtures Implications f o r risk ass e ssm ent inclu d e the f o llo wing:

• Som e m etals act add i tiv ely when they are present together, others act independently of each other, and still others are antagonistic or synergis tic. Such interactions are important aspects of assessi ng exposure and effects.

• Interactions am ong metals within organi sm s m a y occur when they comp ete for binding locations on specific enzym e s or recep tors during the proce sses o f absorption, excretion, or sequestration, or at the target site.

• The presence of and amount of other m e tals are im portant when conducting and interpreting laboratory tests.
1.4.3. Some Metals are Essentia l for Maintain ing Proper Health of Humans, Animals, Plants, and Microorgan isms Implications f o r risk ass e ssm ent inclu d e the f o llo wing:

• Adverse nutritional effects can o ccu r if essentia l m etals ar e n o t available in suf f i cient am ounts. Nutritional deficits can be inherently adverse and can increase the vulnerability of hum a ns and other organi sm s to other stressors, including those associated w ith other m etals.

• Excess am ounts of esse ntia l m etals can resu lt inadverse effects if they o v erwhelm an organism ’s hom eostatic m echanis m s . Such homeostatic con trols do not apply at the point of contact between the organi sm and the environm ental exposure.

• Essentiality thus should be viewed as part of the overa ll dose-response relationship for those m e tals shown to be essential, a nd the shape of this relationship can vary am ong orga nism s. For a given population, “ref erence doses” designed to protect from toxicity of excess should not be se t below doses identified as essential. Essential do ses are typically life-stage and gender specific.
1.4.4. The Environmental Chemistry of Metals St rongly Influences Their Fate and Effects on Human and Ecological Recepto rs Unlike organic chemicals, m e tals are neith er created nor destroyed by biological or chemical processes. Ho wever, thes e processes ca n transform m e tals from one species to another (valence states) and can convert them between inorganic and orga nic form s. Metals also are present in various sizes, f r om s m all particles to large m a sses. Implications for risk assessm ent include the following:

• The for m of the m etal (chemi cal speci es, compound, m a trix, and particle size) influences the m e tal’s b i oaccessib ility, bioavailability, fate, and effects.
1-10

### • The for m of the m e tal is influenced by envi ronm ental properties, such as pH, particle size, m oisture, redox potential, organic m a tter, cation exchange capacity, and acid volatile sulfides.

• Certain form s of m e tals are used for eval uating exposure and effects. For example, the free m e tal ion is used for exposure assessm e n ts based oncomp etitive binding of m e tal to specific sites of action.

• Meta ls attac h ed to sm all airborn e partic le s are of prim ary importance for inhalation exposures, although a few m e tals and m e ta l compounds m a y exist as vapors (e.g., m ercury).

• Inform ation developed on the fate and eff ects of one for m of a m e tal m a y not be directly applicable to other form s.

• Organom etallic form s ha ve different characte ristics from inor ganic m e tals and m e tal compounds, and the sam e general principles and approaches for risk assessm e nt 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 Organi sm’s Ability to Regulate and/or Store the Meta l These processes are often highly dyn amic (e.g., vary acco rdin g to exposure route and concentration, m e tal, and organism ) and thus exert a direct influence on the expression of m e tal toxicity. Implications for risk asses s m ent include the f o llowing:

• Certain m etal com pounds are know n to bioaccumulate in tissues and this bioaccumu lation can b e related to th eir tox i city.

• The latest scientific data on bioaccu mu lation do not currently support th e use of bioconcen tration facto r (BCF) and bio accumu lation factor (BAF) values when applied as generic threshold criteria for the hazard potenti al of inorganic m etals in hum an and ecological risk assessm ent (e .g., for classification as a persistent bioaccumu lative toxic [PBT] chemical).

• Single value BAF/BCFs hold the m ost valu e for site-specific assessm ents when extrapolation across different e xposure conditions is mi nimi z e d.

• For regional and national assessm e nts, BA F/BCFs should be expr essed as a function of m edia chemistry and m etal concentration for particularsp ecies (or c l o sely re lated organism s).

• Trophic transfer can be an important r oute of exposure for m e tals, although biom agnification of inorganic form s of m e tals in food webs is generally not a concern in m etals ass essm ents.

1-11

### • Kinetic-based bioaccumulation m odels (e.g., DY NBAM) have been shown to accurately d escrib e bioaccumu lation resul ting from different exposure ro utes for various m e tals and aquatic organism s and shoul d be considered as alternatives to the BCF/BAF approach when appropriate data are available.

• Many organism s have developed physiologi cal or anatomical m e ans for regulating and/or storing certain m e tals up to certain e xposure levels such that m e tals m a y not be present in organism s in a concentration, for m , or place that can resultin a toxic effect.

• The organ or tissue in w h ich m e tal toxicity occurs m a y differ from the organ or tissue(s) in which the metal b i oaccu mulates and m ay be affected by th e m etal’s kinetics. Target organs m a y differ by species, m a inly owing to differences in absorption, distribution, and excr etion. Effects at the portal of entry to an organism are less dependent on kinetic proce sses internal to an organism.

• Both the exposure route and the form of a m e tal can affect the m e tal’s carcinogenic potential (assessed in the c ontext of hum a n health risk assessm e nt) and its noncancer effects.

• Sensitiv ity to m etals vari es with age, se x, pregna ncy status, n u tritiona l status, and genetics (due to genetic polym orphism s).

1-12

### 2. FRAME W ORK FOR METAL S RISK AS SESSME N T The following discussion addresses issues that are unique to inorganic metals and routinely encountered during the inor ganic 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 assessm ent phases and assessm ent questions. Environm ental chemistry issues and th eir implication in the a ssessm ent of inorgan i c m e tals are also discussed. The chapter is or ganized around the overall risk assessm ent paradigm

. (See Figure 2-1, which b r oadly illu strates the ove ra ll risk asse ssm ent/risk m anagem ent process and identifies som e m e ta ls-specific considerations in the problem for mulation and analysis steps.) An effective risk assessm ent for m e tals w ill account for the un ique aspects of metals that differentiate them from other substances early and throughout the risk assessm ent process. For assessm ents of human health or ecological risks at national, regional, or site-specific scales, the m etals principles can be trans l ated in to sets of assessm ent questions. As appropriate, the risk assessor can us e these ques t ions to m eet the needs of the assessment. The risk assesso r should consider these questions throughout the risk assessm ent process; however, they are especially important in focusing the assessm en t during the Planning and Problem For mulation phase. Suggested assessm e nt questions are given within this Fr am ework for Problem Formulation, Analysis, and Risk Characterizatio n. 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 P R OBLEM FO RMULATI O N

Planning and Problem Formulation are criti cally important for both hum an health and ecological risk assessm ents (U.S. EPA, 2003a , 2000a, 1998a). The concepts em bodied in Planning and Problem Formulation are valuable starting points for any risk assessm ent involving
m e tals. Planning and Problem For mulation provid e an opportunity for ini tial consideration of the m etals’ c h arac teri stic s and the i r c h emistry. Th ese considerations, along with other aspects of the assessment, contribute to the developm ent of a conceptual m odel that conveys the important elem ents of the m etals risk assessm ent. Although Problem For mulation is not explici tly included in the hum an he alth risk
assessm ent (HHRA) paradigm , as it is in the eco logical risk assessm en t guidelines, current practice is to consider m a ny of the issues in th e planning stages that th e as s essor an tic ip ates will be incorpo r at ed later in the HHRA. This is particularly true for m ore complex assessm e nts that

2-1

### Figure 2-1. Risk assessment/risk manageme nt process for meta ls.

2-2

### consider mu ltip le m etals , pathways/routes of expos ure, etc., as is advised in EPA’s Fram e w ork for Cumulative Risk Assessm ent (U.S. EPA, 2003e ). These planning an d scoping activities m a y include

• defining the geographic scale and scope (site, national-scale, etc. ) of the assessm ent • identifying potentially exposed populations and sensitive subpopulations • characterizing exposure pathways and exposure routes (conceptual m odel) • describ i ng h ow exposure will be assessed • determining how hazard and the recepto r’s do se-response will be assessed, and • describ i ng h ow risks wi ll 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 regi onal and national-scale assessments, more general assumptions about th e form of the metal in the environment, deposition pathways, uptake and bioavailabil ity parameters, and sensitive species or subpopulations are useful. These general assu mptions 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 c onsidered in all risk assessments. For metals risk assessment, the risk assessor should consider the following examples of questions that should be consid ered during the planning and scopi ng of the Problem Formulation phases: Transla t ing the Metals As sessme nt Principles into As ses sment Ques tions Tran slatin g the metals p r in ciple s into asse ssm ent que stion s sh ould be stre ssed du ri ng plan ning a nd pro blem formulatio n. This step hel ps en su re tha t the prin ciple s have been a ppro pri ately c o ns ide r ed

.



### Background Concentrations

• How should background (natural and anth ropogenic) levels for m e tals be characterized for the selected sp atial scale of the assessm ent?

• Is ecoregion -specific inform ation availab l e or is the use of state averages , or distri butions , com p atible with the lev el of ecological relevance and certainty required by the risk analysis?

• For ecolog i cal risk assessm e nts, are acclim ation, adaptation, and tolerance data for organism s of concern av ailab l e and are thes e issu es being con sidered ?
2-3

### Mixtures a nd Interactions

• How will interac tions af f

ecting up tak e and syste mic eff ects be conside r e d ?

• W ill issues be conside r e dsuch as re ducti on of re activ ity and incre ase in mobility by organic compounds that for m comp l e xes withm e tals and possible increases in toxic effects of organic com pounds that f o rm lipophilic com p lexes with m etals?

• W ill interactions with o ther m etals a nd with organics (e.g., A s and polycy c lic arom atic hydrocarbons) be addressed? Essentiality

• Are nutritional deficits, which can be i nherently adverse and can increase the vulnerability of hum a ns and other organism s to other stressors, be included in the assessm ent?

• How will both toxicity and def i cienc i es of essential m etals be charac teri ze d? Forms of Metals

• Since enviro nm ental che mistry is a p r im ary factor influencing m etal spec iation and subsequent transport, uptake, and tox i ci ty, how will it be inc l u d ed in the risk assessm ent?

• How will environm ental conditions (e.g., pH and oxidation-reduction reactions) b e addressed to determine m etal speciation and m obility ? Toxicokinetics/Toxicod y namics of Metals

• W h at m etal- related r esp onses are of most concern in th e he alth risk asses s m ent?

• Which sensitive subpopulations should be considered for each m e tal of concern?

• How will biotic and abiotic f acto r s that inf l uenc e the bio a vailability and bioaccumu lation of m etals be incorp o r ated into th e risk assess m ent?

• How will environm ental f actors th at af f

ect m etal speciation and m etabolic capac ity of biota to regulate internal m e tal concentra tions (hom eostatic controls) be accounted for when calculating the bio accumu latio n potential o f m etals ?
2.2. METALS CONCEPTUAL M O DEL The relationships between the sources, expos ure, and effects of m e tals to hum a n and ecological recepto rs are comp lex and often are s p ecific to a particular site, environm ental condition, and recep tor organism

. Because m e tals are natu rally occurrin g substances that undergo extensive biogeochemi cal cycling (i.e., ar e not destroyed but change for m ), transition functions between environm enta l loadings, m e dia concentrati ons, exposed receptors, and the
2-4

### final organism al or ecosystem responses are a ffected by natural processes to a much greater extent than those that occur w ith xenobiotic organic contami nan ts. The a ssessor sh ou ld iden tif y these trans i tion f unctions in the con ce p tual m odel f o r all m etals assessm ents. The generic conceptual model depicted in Fi gure 2-2 shows the in terrelationship betw een the m e tals or m e tal compounds of interest and th e health risk assessm ent process. It is a representation of the actual and potential, direct and indirect rela tionships between stressors in the environm ent and exposed hum a ns (or particular subpopulations) or ecological entities. The conceptual model depicts possible pathways from sources of m e tals and typical ways in which risk is assessed (e.g., on the basis of m e dia c oncentrations, calculated dose, or residues in tissue s). Th is m odel f o llows the same f o rm at as a typic al che mical assess m ent, but it identif ies areas (pri m arily in the tr ansi tion states between environm en tal comp artm ents) where m etalspecific issu es require ad dition al con sideration. For sim p licity, the num erous environ m ental or biological processes that influence the predomin ant route of exposure or the physical/chemi cal properties of the m e tal compounds are not depicted in this model, but su ch processes would be used as inputs to m odels developed f o r specific assessm ents. The bidirectional arrows represent the fact th at the transitio n func tions (models) can be applied in a prospective m a nner (i.e., in a left-to-right direction to determ ine risks associated with a give n load or exposure) or in a retrospective m anner (in a right-to -left direction to determine the load or exposure associated with a predetermi ned level of risk). The latter is usually done fo r generating hum a n and ecological q u ality criteria expres s ed as m edia concentration s . The goals an d scope of a health risk a ssessm ent, in addition to the availab i lity of data,
m ethods, and resources, are am ong the m o st i m portan t factors that determine the extent to which the key prin ciple s spec if ic to m etals (given in Section 1.4) can be incorpo r ated into an
assessm ent. Generally, health risk assessm ent endpoints are selected during the Problem Formulation phase of a risk assessm e n t based on their relevance to risk managem e nt goals, societal values and laws, known adverse effect s of m e tals, and endpoints of i mportance to stakeholders. Risk assessors will incorporate the m etals princ i ples to a les ser extent in
screen ing- le vel asse ssments th an in detailed ri sk assessm ents. Site-sp ecif i c assessm e n ts can account for more m e tal-specific processes (parti cularly environm ental chemistry) than can nationa l- lev el asse ssm ents thatrequ ire gen era liz ation a cross multiple eco regions. Th eref ore, it is
recomm ended that, when appropriate, regional- ornational-level ecological risk assessm ents be subdivided into m etal-related ecoregions, referred to as metalloregions (McLaughlin and Sm olders, 2001), such that pro t ectio n levels, mitigation goals, and ranking results will be appropriate for the suite of species n a turally p r es en t with in e ach type of controlling environm ent. This is directly analogous to the use of ecoregions when establishing water quality criteria (Griffith et al., 1999). T h e Problem Formulation phase of the assessm ent should clearly identify whether a regional approach is be ing used and, if so, how the m e ta lloregions are defined in term s of species composition and envi ronm ental controlling factors.

2-5Media-basedToxicity ModelsDose-based(M8)Toxicity Modelsdietsoil, seds.,air(M7)fur, feather etc.soil, seds., airother endpointsground watersoil, seds., airgrowth, repro.surface waterkidney, gillground waterground waterliver, lung,mortalitysurface watersurface waterRiskMetalMetalMedia-based(Toxicity)Residue-basedBioaccumulationMetalFate & TransportDistributionExposureExposureToxicity Modelsto/TK ModelsDistributionModelsLoadingsinModelsto MetalOrganism /in Mediato MediaTissuesIndividual(M2)(e.g., mg/kg/d)(M1)(M3)(M4)Population,Dietary ExposureHabitat, EcosystemModelsModels(M6)(M9)ecosystem2oconsumerscommunity1oconsumerspopulationproducersRisks atMetalHigherBioaccumulationDistributionLevels ofFood Web ModelinBiologicalDietOrganization(M5)Figure 2-2.Generic conceptual model for metals riskassessment.



### 2-6

### This concept of regional-based e colo g ical assess m ents is sig n if icantly le ss important in human health assessm ents. In th ese assessm ents, th e environm ental controlling factors (pH, water hardness, etc.) m a y be important determin ants in exposure calculations for dietary or drinking water exposures. However, to our know ledge , hum a ns have not adapted to particular areas of m e tal enrichm e nt or impoverishm e nt but, rather, choose to live in all environm ents. Therefore, the differences in hum an sensitivity that should be considered are not geospatially correlated. Rather, the assessor should strive to identify pot entially sensitive subpopulations, such as the very young or the elderly, subpopulations withg enetic pred is position s to m etal sensitivity (e.g., W ilson’s diseas e), 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 th es e groups will be given addition al sc rutiny and s e parate risk analyses, su ch that results will be applicable onl y to the general population. Figure 2-2 identifies areas in th e con ceptual m o del that stand out as m e tal-specific iss u es as the transitions between environm ental loadings , m e dia concentrations, exposure receptors, and the final o r g anism al or ecosystem ris k. Because m etals are n aturally -occ urring sub stances with which organism s have evolved, it is particular ly important to incorpor ate the natu ral p r ocesse s that affect metal m obility, speci ation, biogeochemi cal cycling, a nd sequestration into the health risk assessm ent. These m a y differ in detail s or approach, depending on the environm ent of concern (water, land, and air), th e final receptor organism s (hum an s, anim als, and plants), and the m a nagem e nt goal (i.e., whether the m a nage ment goal is the health of individuals or the
m a intenance of populations or communities). Howe ver, the sam e basic concepts always arise, regardless of the assessm e nt context. The conceptual m odel identif ies the f o llowing is s u es, indicate s the poin twithin th e h ealth risk as sessm ent process where they occur, and h e lps direct the rem a inder of the health risk
assessm ent.

• M1: Fate a nd transport models. T he partitioning and biogeochemical cycling of m e tals into the various environm ental m e dia from the loading source depend on the physical properties of the ini tial form of the m a terial and the particular chemistry of the receiv i n g environm ent. Fate andtransp ort m odels are useful for estim a ting m e tal speciation, transition ki n e tics, partiti oning, depos ition, and po tential resus p ension within th e c ontext of en vironm ental le vels of the m e tal and other inorganic substances. These can be very detailed fo r site-s pecific as ses sm ents, or they can provide a potential range of pr ocesses that mi ght occur over large, regional scales for assessm ents of a m ore generic nature (e.g., criteria developm ent or ranking schem e s). Reviews by Paquin et al. (2003), Allen ( 2002), and EPA (1997a) include up-to-date inf o rm ation with regard to the availab ility of m odels appropriate for use in evaluating fate and transport of m e tals in aquatic e nvironm ents (see Chapter 3 on environm ental chemistry).

2-7

### • M2: Media -based exposure models. Media-specific exposure m odels are m a the m atical functions used to calculate the exposure of the organism to m e tals directly from abiotic m e dia (i.e., excluding the food web). Estim ating the uptake of m etals from environm ental m edia into biot a follows m a ny of the sam e processes used for organic substances, such as understandi ng dietary preferences, ingestion rates, inhalation rates, and m ovem e nt patterns. Of particular concern with m e tals health risk as sessm ents is accou nting for the diffe ring bioavailability of m etal species to organism s from different environm ental m edia. Exposure to existing environm ental levels of m e tals is another issue of cons iderable importance in this m odeling step. Exposure assessm ent issues are considered se parately for hum a n health, aquatic, and terr estria l re ceptors (se e Chap ters 3, 4, and 5, respectively).

• M3: Bioacc u mulation and toxicokinetic (TK) models. Many organic substances require m e tabolic activation to becom e t oxic or, conversely, to be detoxified and excreted. Metals do not. Metals m a y for m compl e xes with proteins or other carrier molecules f or distribution to target organs or for sequestration and excretion. Their bioaccumu lation is tissue-specific (e. g., cad mi um [Cd] in kidneys). The natural occurr ence of m etals ha s led to the d evel opm ent of specific m echanism s for uptake, m e tabolism , distribution/storage, and excr etion of m e tals by organism s. These processes can im pact the use and in terp retation of bioaccumu lation data an d the toxicity of bioaccumu lated m e tal.

• M4: Resid u e-based to xicity mode ls. If risk to the organism (s) of conc ern is to be based on an estim ate of internal dose, th en 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) m o dels. Becau se of the pro cesses d i scu ssed in th e p r evious paragraph, th is can be particularly challengi ng for inorganic m e tals. Meta l speciation in the exposure m atrix can especially influence this relationship because up take and organ distribution kinetics are like ly to differ. When ava ilable, critical body residues (CBRs) can be used to reduce uncertainties in health risk assessm e nts because they account for site-specific bioa vailability and mu ltiple expos ure pathways. However, CBRs for metals can vary widely depe nding on exposure pathway (food vs. water), rate of accu mulation relative to the detoxi ficatio n and seques tration pro c esses, and for m of bioa ccumu lated m e tal. Establishm ent of a valid residue-response relationship is critical for successful applicati on of CBRs (see Sections 5.3 and 6.3).

• M5: Bioaccumulation/food w e b model. Movem e nt of metals through the food web is complicated by factors of bioaccessibility, bio a vailability, essentiality, regulation o f m e tals (uptake and internal distribution) , detoxification, and storage as well as accumu latio n and the natural adaptiv e capacity of organism s. While the ab ility to quantitatively address all these facto r s m a y be limited at present, the assessor should at least qualitatively addr ess the po ten tial im pacts. Bioaccu mulation an d trophic transfer of m e tals does occur. Howeve r, biom agnification (i.e., increases in concentratio n through multip le levels of the food web) is rare, with the ex ception of certain organom e tallic compounds, such as m e thyl m e rcury, that can biom agnify m any orders of m agnitude in the aquatic food chain. Discussions of m e thods for estim ating b i oaccumu lation in aquatic and te rrestrial food webs are found in Sections 5.2.5.3 and 6.2.5.2, respectively. 2-8

### • M6: Dieta r y exposure models. Th e assesso r must carefully consider the bioavailability of m e tals from food item s inm ode ls that estimate dietary exposure to m e tals. In ecological risk assessm ents, the wide variation in feeding m odes and digestive ph ysiology across species limits th e ability to m ake generalizations with m e tals. Generalizations requi re knowledge of dietary pref erences, trophic structure of the community, and ingestion and absorp tion rates. In hum an health risk assessm ents, the assessor should consider regional, social, a nd religious dietary preferences. Although this node of the conceptualm odel differs very little from risk assessm ent approaches for organic substances , s o m e m etal-s p ecific generalities about the relative importance of exposure pathways can be applied to focus (and simplify) the process. For exam ple, the high es t accu mu lation of m e tals in plan ts generally occurs in the roots, and, except for h ype raccumu lato r species, m o st plant trophic transfer rates can be assum e d to be <1. Ther efore, direct tox i city to herbivores is less likely than f o r insectivores or from other dietary pathways, and risk to hum a ns from most fruits and vegetables (except roots or green, leafy vegetables) is low. On the other hand, plants are quite sensitive to so m e m e t a ls and m a y die before achieving levels high enough to be toxic to anim als, thereby affecting them indirectly through reduction in food availability. A discussion of dietary exposure assessm ent issues is found in Sections 5.2 an d 6.2 for aquatic and terrestrial ecolo gical recep tors, respectively, and in Section 4.2.5.2 for hum ans.

• M7: Exposure-based toxicity mod e l. Calculation of an external dose (oral intake, gill b i nding, etc.) f o r co mparison with toxicity thresholds m ay depend on inf o rm ation about relative bioavailability (RBA), speci ation of the m e tal or m e tal salt, dietary preferences and rates, en vironm ental c oncentrations, essentiality, and m e tal interactions. Toxicity threshold consid erations should be based on compar able inform ation, such as appropriate m e tal species in exposure media, similarly acclim ated or adapted organism s, similar exposure routes, and appropriate com b inations of essentia l m etals. Chemi cal equilibrium m o dels such as MINTEQA2 (Brown and Allison, 198 7) m a y be useful for characterizing the species o f m e tal that is present in particular media, m aki ng exposure and effect comparisons more comp arable. This form s the basis of the biotic ligand m odel (BLM) approach (Di Toro et al., 2001; Santor e et al., 2001; Paquin et al., 1999) to defining acute aquatic toxicity.

• M8: Media-based toxicity model. This health risk assessm ent m odel compares environm ental concentrations with organism response func tions without calculating a body burdenor internal dose. It is u sed m o re frequently for aquatic and s o il-dwelling organism s, less frequently for wildlif e, and very infrequently for hum an health assessm ents. Consideration of RBA, trophi c transfer rates, dietary preferences, existing environm ental concentrations, and organism adaptations is important for a m e tals assessm e nt.

• M9: Population, habitat, ecosystem models. Assessors w ho carry out Ecological risk assessm ents often ask questions relate d to population growth, habitat change, or ecosystem functions in addition to questions re lated to risks to individual organism s. Most of the models and approaches are similar f o r both m e tal and organic substances. However, metals and other inorganic substances are am ong the funda m e ntal determinants and delimi ters of ecoregions (in conjunction with cl im ate, elevation, and 2-9

### day length associated with latitud e ). Ther efore, knowledge of backgroundlevels and the adaptation of organism s to differing m etal le vels ar e ess e ntia l in deve loping appropriate risk factors for naturally occurring species.
In summ ary, the conceptual m odel lays out a series of working hypotheses about how the
m e tal(s) of concern mi ght m ove through the enviro nm ent to cause adverse effects in hum a ns or ecological system s. These hypotheses are examin ed through data analyses, m odels, or other predic tive to ols to de termine the probability and m a gnitude of the occurrence of unwanted effects. The approaches used to acco mplish th is assessem e nt are discussed in general within various Agency risk assessm ent guidance documents.

2.3. ASSESSMENT PHASE The assessm ent phase of a heal th risk assessm ent is the process of estim ating exposure and understanding the dose-response re lationship between biota and th e chemical(s) of interest. The additional m e tals-specific factors should be considered during this phase. As with any
assessm ent, at the beginning of the Analysis phase, the as sessor shou ld critica lly e x amine the Bioaccessibi lit y , Bioa v a ilabilit y , and 

### data and m odels to ensure that they are Bioacc umulation 

### appropriate to the lev el o f detail and site -spec i f i c, Bioacc essibi lity refers to the amou nt of 

### regional, or national app lication of th e assessm ent envi r onm enta lly available m e tal that actually intera cts with the orga nism’ s co ntact surf ace (e.g., 

### results. Mo st of the as s essm ent questions in th is membrane ) a nd is pote n tia lly available for 

### chapter are directed tow ard assisting the assessor absorptio n (o r adsorption if bioactive up on c ontac t). Environm entall y

avail able m e tal is the 

### with the co llection of the appropriate inf o rm ation total amount of metal that is availabl e for physi cal, chemi c al, an d biologi cal m odifying influe nce s 

### to address metal-specif ic issues for conducting (e.g., fate and transpo rt) a nd is not seq ueste red in 

### either exposure or effects characterizations. an enviro n me ntal matrix. Bioavailability of metals is the extent to whic h bioa ccessible metals ab so rb onto, or into , and 

### 2.3.1. Bioavailability across biol ogi cal mem b ra n e s of organi sms, expre s sed a s a fraction of the total amou nt of 

### The bioavailability of m e tals and, metal the org anism i s proximately exposed to (at the sorption surface) d u rin g a given time and 

### consequently, the associated risk vary widely unde r define d conditio ns. 

### accord ing to the physical, chemical, and Bioacc umulation of meta ls is the net accumul a tion of a metal in the tissu e of intere st or 

### biological conditions under which an organism is the whol e org anism that re sults from all 

### exposed. To the extent that available data and envi r onm enta l expo su re m edia , incl udin g air, water, solid p hases (i.e., soil, sedime nt), and diet, 

### m e thods allow, the assessor should explicitly and that rep r ese n ts a net mass bala n ce betwe en 

### incorporate factors th at influence the uptake and el imination of the metal (SA B , 2006). Biocon cen tr ation is the n e t accumulati on of 

### bioavailability of a m e tal into the health risk metal in an organi sm re sult ing from direct uptake


### assessm ent. In situations where data or m odels from water only, such as through gill membranes or othe r external su rfaces. 

### are insuf f i cient to addr es s bioavailab i lity rigorously, the assump tions m a de regarding
2-10

### bioavailability should be clearly detailed in the healthrisk assessm ent as should the associated impact on results. Although bioavailability m a y be a defined m easurem ent when considered in certain verteb rate anim als where m e tal uptake is directly a function of the concentr ation of m etal in the diet, it is not as sim p le in m a ny other aquatic and terrestrial organism s where food consumption is difficult to m easure and where metals ar e present in the surrounding environm ent and available for uptake via nondietary pa thways. In this case, as di scussed in Meyer (2002), m e tal bioavailability m a y be more of a conceptual term and not a precisely m easured param e ter. Environmental ava ilability ref ers to the ability of a m e tal to in tera ct with o ther environm ental m a trices and undergo various fate and transport processe s. Environm entally available m etal is not s eq u ester ed in an environm ental m atrix, and it represents the total pool of
m etal in a sy stem that is potentially bioavailable at a particular tim e and under a particular set of environm ental conditions (i.e., able to contact or enter into an organism ). Environm ental availability is specific to the existing envi ronm ental conditions and is a dynami c property, changing with environmental conditions. The bioaccessible fraction of m etal is the portion (fraction or percentage) of e nvironmentally available m e tal th at actually in teracts at the organism ’s contact surf ace and is p o tentially av ailab l e for absorption o radsorption (if bioactive upon contact) by the organism

. The bioaccessibility, bio a vailab ility, and bioaccu mulation prope rties of inorganic m e tals in soil, sedim e nts, and aquatic system s are in terrelated and abiotic (e.g., organic carbon) and biotic (e.g., uptake and m e tabolism )

. Modifying factors determine the amount of an inorganic
m e tal that in teracts at bio l ogical surfaces (e.g., hu m an digestive system , at the g ill, gu t, or root tip ep ithe liu m ) and that binds to andis absorb edacross these m e mbranes. A m ajor challeng e is to consis ten t ly and accu rately m easure quantit ativ e differences in bioav a ilability between multiple f o rm s of inorganic m etals inthe environ m ent. The bioavailability issue paper authors (M cGeer et al., 2004) provided EP A with som e practica l , sta ndard, and d e f e nsible rec o mm endations on conce p ts, te rm s, and def i nitio ns that can serve as a paradigm for s tudying inorganic m e tals and their bioavailability. Figure 2-3 presents a conceptual fram ework along with further di scussion of m e tals bioavailability and bioaccumu lation.

2.3.2. Exposure Characteriz ation Exposure characterization desc ribes potential or actual c ontact or co-occurrence of stressors with recep tors (U.S. EPA, 1998a). Meta l factors incorporated in to this po rtio n of the health risk assessm ent include eco sys t em and recepto r characteris tics that a ffect the movem e nt of
m e tals in the environm ent including atm o spheri c deposition, their up take and accumu lation in hum a ns and other biota, and di stribution into target organs. S pecific assessm e nt questions include the f o llowing: 2-11Total 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 c i a t i o n Uptake efficiency

Dietary intake Uptake efficiency

Bioaccumulated Metal Uptake efficiency

a d s o r b e d Bioavailabilit

a b s o r b e d Membrane

uptake

Detoxification and Storage

Bioaccumulation of metal

Excretion Benign accumulation

Essentiality

Predation ForagingRoot 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 e valuating 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 bio accumulation or trophic transfer. Source: McGeer et al. (2004).

2-12

### Background Concentrations

• What data sources are used to estim ate background (natural and anthropogenic) concentratio ns?

• What are the ranges of background concentr ations and how do they vary spatially?

• What degree of certainty exists in es tim a tes of bac kground concentrations? Fate and Transport

• W h at enviro nm ental tran sport and ai r depos ition models will be used and what are their assumptions, limitations, and uncertainties?

• How will the environm ental ch emistry (in ai r, w ater, and soils) of m etals be addressed?

• What deposition scale (local, regional, or national) is important f o r the m etal and recep tors being consid ered?

• What m e teorological factors impacting the fate and transport of m e tals should be considered in the health risk as sessm ent? Mixtures a nd Interactions

• Is exposure to m e tal mi xtures being incorpor ated into the exposure assessm e nt? If so, how is it being addressed? If not, what is the rationale for not addressing m e tal mi xtures?

• What eviden ce exis ts to indicate expo sure to the m e tal(s) of concern is affected by m etal mi xtures in the as sessm ent? Essentiality

• For essential m e tals, will exposure co ncentr ation s exceed th e nutritional requirem e nts (recomm ended dietary allowance [RDA])?

• How do the nutritional requirem e nts vary across species and populations in the assessm ent? Forms of Metals

• What for m s (chemi cal sp ecies ) of m e tals are lik e l y to occur at the site( s) o f interes t ?

• What biogeochemi cal speciation and transf orm a tion processes are relevant for the assessm ent?

• How mi ght these biogeochemi cal processe s im pact exposure assessm ent for the m e tal(s) of concern?

2-13

### • What transport and fate models are rele vant for the environm ent and m e tals of concern ? Toxicokinetics/Toxicod y namics of Metals

• What enviro nm ental factors have the most influence on the b i oavailability of the m e tals of concern?

• What m e thods will be us ed to addres s bioavailab i lity in the as sessm ent?

• How is bioaccumu lation being asses sed or pred icted ?

• To what extent are bioaccumulation predic tions being extrap olated acros s species, exposure concentrations, locations , or environm ental conditions?

• What are the key assump tions being us ed to address bioaccumulation and bioavailability and how accurate are these as sumptions ? The objective is to produce a comp lete picture of how, when, and where exposure occurs or has occurred by evaluating sources and releas es, the distribution of the stres s or in th e environm ent, and the extent and pattern of contact or co-occurrence with hum a ns or ecologically relevant biota. The m e tal-specific exposure fact ors discussed in this fram e work contribute to the exposure characterization, but addi tional issues that are generall y applied to all health risk

assessm ents also should be considered (although th ey are not specifically discussed here). For the exposure profile to b e useful, it s hould be co mparable with the s tre ss or-response relationsh i p generated in the eff ects characterization.

2.3.3. Characteriz a tion of Effects/Haz a rd Analysis To characterize effects or adverse responses to metals, the risk assessor should describe how the effects are elicited, link them to the hum a n populations at greatest risk and/or the ecological assessm ent endpoints, and evaluate how th ey change with varying exposure levels. It is particularly important, especially f o r inorga nic m e tals, to confirm that the conditions under which the ex posure occu rs are consis tent with tho se of the con ceptua l m o del. This will ensure that the correct m e tal species is evaluated for its effects on the popul ations (including the vulnerable subpopulation) or endpoints of concern, or that appropriate m odels are used for extrapolating responses am ong meta l species, biota (laboratory to field, or test species to hum a ns), or for varying environm ents (e.g., m e talloregions). Assessm ent questions regarding
m e tal-specif i c factors for effects analyses or hazard asses sm e nts inc l ude the following :

2-14

### Background Concentrations

• What is the relationship between envi ronm ental (natural and anthropogenic) concentrations and toxicologically relevant m e tal concentrations?

• For ecological risk assessm e nts, how are ac clim ation, adaptation, and tolerance issues being addres sed in th e effects analy sis?

• In hum an he alth assessm ents, have concentra tions in locally grown or harvested foods been taken into account when estim ating elevated m etal exp o sures or estim ating relative bioavailability of m e tals in foods, soil, or water?

• How represe n tative ar e the toxic ity test cond itio ns of the environm ents being assessed ?
Mixtures a nd Interactions

• Are toxico lo gical effects of m etal mi xtur es being incorporated in the effects assessm ent? If so, how are they being addre ssed? If not, what is the rationale for not addressing the toxicity of m e tal mi xt ures?

• For particular mi xtures of inorganic m e tals , to what degree are their com bined effects additive, antagonistic, or synergistic?

• Is mi mi cry (com petitive interactions am ong che mically similar m etals/m etalloids ) important inthe asse ssment?

• For site-specific ass e ss m e nts, what evidence exists to indic a te the tox i city of the m e tal(s) of concern is affected by the presence of other m e tals? Essentiality

• For essential m e tals, are nutritional re quirem e nts known (e.g., RDA for hum a ns)?

• What is the range between concentrations required nutritionally a nd those associated with toxicity reference values (e.g., refe rence concentration [ R fC], Ambi ent W a ter Quality Criteria [ A W Q C] ) or adver se ef f

ect lev els used in the risk a sses sm ent?

• Are nutritio nal deficits that can increase the vulnerability of hum a ns and other organism s to other stress ors being addressed?
Forms of Metals

• Which for m s of the m e t a ls are m ost toxicologically relevant?

• How toxicologically co mparable are the fo rm s of m e tals used in the ef fects and exposure assessm ent?

• How mi ght assumptions regarding the toxic ity of different m e tal form s i mpact the effects asses sm ent? How accurate are thes e assu mptions ? 2-15

### • How will the atm o spheric chemistry , transpo rt, and depositio n of m etals be address e d in the assessm e nt?

• What m eteorological factors im pact the fate and transport of m e tals?
Toxicokin e tics/Toxicodynamics of Metals

• How does toxicity vary f or different m e tal for m s found (or likely to be found) in the assessm ent?

• Which of these are m ost important and wh ich are incorpo r atedinto the e f f

ects assessm ent?

• How are absorption, distribution, m e tabolism , and excretion addr essed for individual m e tals or mi xtures of concern? Effects analysis results are summ arized ina stressor-response profile. The analysis addresses the plausibility that effects m a y occu r or are occurring as a result of exposure to the
m e tal(s) of concern, and that li nkages between m e asured effects and assessm e nt endpoints can be
m a de (this is especially important for ecological risk assessm e n ts). Many of these step s in effects assessm e nts are not unique to inorganic metals, and so they are not addressed s pecifically in this Fr amework. Although the prediction of toxicity due to dietary exposure to i norganic metals is
comp licated by wide variation in th e bioavailability and toxicity of m etals, it is a f actor that risk
assesso rs sh ould consid er in m etals assessm ents. Direct approaches to accomp lish this includ e quantifying the bioav a ilable fractionof bioaccu mu lated m etals in consu m ers (e.g., analysis of tissue f r a ctions such as cytosolic m etals ) and de termining m etal sp eciatio n in the m edia of concern (water, soil, or air). Comparisons of m e dia values can then be made to toxicity reference values using the sam e m e tal species. L acking such inform ation, or for higher tier
assessm ents, bioassay m e thods with field-co llected m e dia offer another w a y to assess bioavailability, although other than lead exposure in juvenile swine, such m e thods have not been widely standardized.

2.4. RISK CHARACTERI Z ATION Risk Characterization is the final phase of the health risk assessm e nt and is the culmination of the Planning, Problem Formulation, and Analysis of predicted or observed adverse effects. It com b ines th e resu lts of th e exposure assessm e nt with infor m ation on stressorresponse profiles to estim a te the likelihood of e ffects of specified m a gn itude(s). The risk
assesso r sho u ld describe available lin es of eviden ce and conduct (and report) an uncertainty analysis. Conclusions presented in the Risk Ch aracterization should provid e clear inform ation to risk m a nagers that is useful for decision m aki ng. There are no m e tal- specific m e thods in the
2-16

### Risk Characterization, other than revisiting the m etal factors descri b e d ab ove to verify that they
were acco rd ed proper co nsideration during th e An alysis. Ho wever, because m e tal as sessm ents are dependent on specific attribut es of environmental chemistry a nd biological responses related to the natura l occur r ence of m etals, it is particu l arly im portant that the Ris k Characte r ization specif y the c onditions, lo cation s , and tim e-f r ame within whic h the asse ss m ent results are applic able. For risk assessm ents conducted for regi onal or national assessm e nts, criteria developm ent, or ranking purposes, it should be acknowledged that re sults will be based on organism s and soil types that result in greatest bioavailability and sensi tivity. Care should be taken, however, that the organism -environm ent com b inations assessed are , in f act, co mpatible with rea l -wo r ld cond itio ns. Relevan tasses sm ent questions in clude th e f o llowing:
Background Concentrations

• What assumptions are made regarding background (natural and anthropogenic) concentratio ns in characterizing m e tal risk s ?

• How sensitive are the risk assessm ent results to the pres ence of background concentrations (i.e., are background concentr ations a m a jor or mi nor component of the ri sk es tim ate)?

• Have m e tals with generally high background concentrations (e.g., aluminum (Al) and iron (Fe ) in soil) b een ap propriate ly c onsidered inecological assessm ents?
Mixtures a nd Interactions

• How sensitive are the risk assessm ent resu lts to assumptions regarding exposure and effects of metal mi xtures?

• To what extent do the m e thods and assumpti ons regarding the exposure and effects of m etal mi xtures introduce intentional or unintentional bias in the Risk Characterization ? Essentiality

• How do risk assessm ent results comp are to levels required to m a intain nutritional health ?

• How sensitive are the risk assessm ent resu lts to m e thods and assumptions used to address e sse ntia lity ?

• To what extent do the m e thods and assumpti ons regarding the exposure and effects of essential m e tals introduce intentional or unintentional bias in the Risk Characterization ?

2-17

### Forms of Metals

• How sensitive are the risk assessm ent resu lts to m e thods and assumptions used to address the different m e tal form s?

• To what extent do the m e thods and assumpti ons regarding the exposure and effects of m etal f o rm s introdu ce in tentiona l or u n inten tiona l bias in the ri sk assessm ent?
Toxicokinetics/Toxicod y namics of Metals

• How sensitive are the risk assessm ent resu lts to m e thods and assumptions regarding factors affecting bio a vailability and b i oaccumulation of the metal?

• To what extent do the m e thods and assu mptions regarding the factors affecting bioavailability and bioaccumulation of m etals introduce intentional or unintentional bias in the risk assessm ent?

2-18

### 3. ENVIRONMENT A L CHEM IS TRY, TRANSPORT AND F A T E

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 th is chapter in the context of risk assessment. Because the behavior of metals defies simple generalities, underst anding the chemistry of the particular metal and the environment of concern is necessary. However, the f actors that control metal chemistry and the environmenta l characteristics used to produce estimates of metal fate and effects can be generalized. Chemical Sp ecies The Intern ational Union of Pure an d App lied Ch emist ry (IUPAC) defi nes chemi c al spe c ie s a s ch emical com p ound s that differ in isotopi c comp osition, confo r matio n , oxidation or electroni c sta t e, or that in the nature of thei r com plexed or covale ntly bound sub s titue n ts, can b e reg a rded a s distin ct chemi c al s p ec ies

.



### In environm ental chemistry, the phase the species occurs in (gas, liquid, aqueous solution, mi neral, o radsorbed on an interface betw een phases) generally is also in clud ed in a
complete definition. In the context of the enviro nm ental chemistry of m e tals, chemists speak of
a m etal species as a “specific form of an elem ent defined as to isot opic composition, electronic or oxidation state, comp lex or m o lecular structure” and phase (Templeton et al., 2000). Meta l spec iation gr eatly determines the be havior and toxicity of m e tals in the environm ent. Speciationrefers to the occurrence of a m e tal in a variety of chemical form s. These form s m ay includ e free m etal ions, m etal complexes dissolved in solution and sorbed on solid su rfaces, and m e tal species that have been c oprecip itate d in m ajor m etal solids or thatoccur in the i r own solids. The speci ation o f a m etal af fects no t only its tox i c ity but also its volatilization, photolys is, sorption, atmospheric deposition, ac id/base equilibria, polymerization,
complexation, electron-transfer reactions, solubi lity and precipitation equilibria, mi crobial transform a tions, and diffusiv ity (Bodek et al., 1988). The following sections address the applicati on of hard and soft acid and base (HSAB) concepts to m e tal behavior, including the form a tion of m etal com p lexes, and the im portance of pH and oxidation-reduction reacti ons to m etal m o bility. Thes e sections also discuss the occurrence and interactio ns of the m e tals of concern in natu ra l m e dia (inc l uding surface and ground waters, soils and aquatic sedim e nts, and th e atm osphere). Metal sorption behavior, aging in soils, m e tal dissolution and transf orm a tion and transfer to plants, and m e thods of determining
m etal speciation in so ils and se dim ents are im portant topics co nsidered in these sec tion s.

3-1

### 3.1.1. Hard and Soft Acids and Bases: The Stability o f Complexes Comp lexes are form ed between m etals (acids) and ligands (bases), both in solution Hard a nd Soft Acids and Bas es 

### and at th e su rfaces of mi nerals and o r ganism s. Hard a c ids a nd hard bas e s . Complex e s 

### The toxic reaction of organism s to metals can formed b e twe en divalent h a rd a c id catio n s an d monovale n t or divalent ha rd bases a r e i onic a nd 

### be dire ctly r elated to th e nature of th e m etal relatively wea kand a r e ofte n termed “ion pairs. ” C o mp le xe s for m e d be tw een Be 2+ or trivalent 

### comp lexes form ed in solution and at the hard a c id s an d hard b ases tend to be ion i c and 

### surface of th e organism

. relatively strong. Soft acid s

a nd soft base s. Relatively s t rong 

### A useful concept that helps to explain covale nt bon ds are forme d in compl e xes 

### the strength of m etal complexing and m e tal betwe en soft and bo rde r lin e soft acid ca tions and soft base s
. Ligand -bin ding site s on the 

### toxicity is that of HSAB, which was external o rinternal su rface s of org anism s are 

### introduced by Pearson (1973)

. In this concept, often soft base and thu s bo nd stro ngly wi th soft and bo rde r lin e soft acid ca tions. 

### m etal cation s are Lewis acids and lig ands are Lewis bases, with the m etal c ation an d 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 m obile and easily m oved. Soft species pref er to participate in covalent bonding. Hard species are comp aratively rigid and nondefor m able, have lo w polarizability, ho ld their el ectrons firmly, and prefer to participate in ionic bonds in com p lex form ation (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 comp le xes. In contrast, the Ligan ds a r e simply anion s or mole cul es that form in co ordi nation compl e xes with met a l ions. 

### bonds for m ed between hard-soft or soft-hard Dep endin g o n wheth er a li gand sha r e s one, two, 

### acids and b a ses are weak , such that th eir three, or m o re electron pai rs with metals, it is calle d a mon o -, bi-, tri-, or multidentate l i gand. 

### complexes tend to be rare. Table 3-1 For liga n d s compo s ed of more tha n on e atom, 

### summarizes hard and soft acid and base the atom that dire ctly partici pates in meta l-ligan d bindin g is called the don or or ligan d. 

### relationships for the m e tals of concern. Hard metals (hard acids), which are the least toxic, preferentially bind with hard bases that contain oxygen, forming weaker bonds with soft nitroge n and sulfur species. The strength of binding between hard m e tals and hard ligands is usua lly a function of pH. Many of the hard m etals are m acronutrients. Soft m etals (acids) bind preferentia lly with soft S and N ligands, forming weaker bonds with hard base speci es such as hydroxide and sulfate. Soft and borderline m etals, and Mn 2+ 

### , which is hard, form bonds of decr easing strength with soft ligands such as sulfide, generally in the following order: Pb 2+ 

### >Cu 2+ 

### >Cd 2+ 

### >Co 2+ 

### – Fe 2+ 

### >Ni 2+ 

### > Zn 2+ 

### > Mn 2+



### . The tendency of m e tals to bind to soft lig ands or to organic s ubs trates (which are

3-2

### Table 3-1. Hard and soft acids (met al cations) and bases (ligands)
Hard acids Al 3+ 

### , Ba 2+ 

### , Be 2+ 

### , Co 3+, 

### , Cr 3+ 

### , Fe 3+ 

### , Mn 2+ 

### , Sr 2+ 

### , U 4+


### , UO 2 2+ 

### , VO 2+ 

### Borderline acids (between hard and soft) Co 2+ 

### , Cu 2+ 

### , Fe 2+


### , Ni 2+ 

### , Pb 2+ 

### , Zn 2+ 

### Soft acids Ag + 

### , Cd 2+ 

### , Cu + 

### , Hg 2+ 

### , Hg + 

### , CH 3 

### Hg + 

### , Tl 3+ 

### , Tl + 

### Hard bases F - 

### , H 2 

### O, oxyanions: OH - 

### , SO 4 2- 

### , CO 3 2- 

### , HCO 3 - 

### , C 2 

### O 4 2- 

### , CrO 4 2- 

### , MoO 4 2- 

### H n 

### PO 4 n- 3 

### , H n 

### AsO 4 n- 3


### , SeO 4 2- 

### , H 2 

### VO 4 - 

### , NH 3 

### , RNH 2 

### , N 2 

### H 4 

### , ROH, RO ­ 

### , R 2 

### O, CH 3 

### CO O - 

### , etc. Borderline bases (between hard and soft) Cl - 

### , Br - 

### , NO 2 - 

### , SO 3 2- 

### , H n 

### AsO 3 n- 3 

### , C 6 

### H 5 

### NH 2 

### , C 5 

### H 5 

### N, N 3 

### -, N 2 

### Soft bases I - 

### , HS - 

### , S 2- 

### , CN - 

### , SCN - 

### , Se 2- 

### , S 2 

### O 3 2- 

### , -SH, -SCH 3 

### , ­ NH 2 

### , R-, C 2 

### H 4 

### , C 6 

### H 6 

### , RNC, CO, R 3 

### P, (RO) 3 

### P, R 3 

### As, R 2 

### S, 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 m etals (sof t ac ids), f o llowe d by the har dm etals (hard a cids ) , typica lly in the orde r Pb 2+ 

### >Cu 2+ 

### >Cd 2+ 

### >Zn 2+ 

### >Ca 2+ 

### >Mg 2+ 

### >> Na + 

### (Pickering, 1986). The tendency of m e tals to for m solid phases, suchas sulfides in sedim e nts, is also related to their HSAB qualities. For example, extrem el y insoluble m e tal sulfides are form ed in anoxic sedim e nts by soft acid m e tal cations, such as H g 2+ 

### (log K sp 

### = –57.25) or Ag + 

### (log K sp 

### = –49.7), whereas borderline hard and hard m e tal cations such as Mn 2+ 

### (log K sp 

### = –19.25) or Fe 2+ 

### (log K sp 

### = –22.39) form slightly more solubl e, although still highly insoluble, m e tal sulfides. Solubility products for all sulfides except Ag 2 

### S 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, m e ta ls and m etalloids c an exis t in the environm ent in s evera l valence form s. They can also exist as or ganom etallic compounds. Organom e tallic compounds
are compounds that have a m e tal/m e t alloid -carbon bond. The bonds in organom etallic
compounds are generally covalent and betweensoft acid m e tals and soft ligands. Metal/m etalloid transform a tion processes, such as m e tal m e thylation, occur through interactions with other chemicals and biota in the environm en t. Cycling and distribu tion of organom e tallic
compounds between terrestrial, aquatic, and atm o spheric phases m a y be physically, chemically,

3-3

### or biologically m e diated. Table 3-2 lists examples of som e commonly occurring, environm entally stable organom etallic compounds.

Table 3-2. Examples of organometallic compounds
Metal/metalloid Organ ometallic compounds AS Methylarsenic acid, dimethyl arsenic acid, trim ethyl arsine, trim ethylarsine oxide Pb Tetram ethyl/ethy l lead, trim ethyl/e thyl lead, dimethyl/ethyl lead Hg Methyl m ercury, dim ethyl m ercury Se Di m e thyl selenide, dim e thyl diselenide, seleno-ami no acids Sn Tributyltin, bis(tr ibutyltin) oxide

As discussed in this section, organo m e tallic m ethylation anddem ethylation rates are influenced by both the speciation and bioavailabili ty of the m etal, the microb ial community, and a large number of environm ental factors, m a ny of which are inte rre lated. Sulf ide andorganic
m atter are important environm ental variab les that significantly affect m e thylation; however, their effect on m e thylation/dem e thyl ation is poorly understood. T he dominant variables differ a m ong locations and between seasons, although it is clear th at m e thylation is predominantly a biologically m e diated process. Methylation/dem e thyl ation rates ar e stro ngly influenced by
m etal/m etalloid spec iatio n and bioavailab ility. Environm ental m ethyl-m etal con centrations ref l ect the ne t m ethylation rathe rthan s i mp le rates of m e thyl-m etal synthesis. T able 3-3 presen ts m e tals that are involve d in abiotic or biotic
m e thylation/dem e thylation processes. W ith the e x ception of As and Se, the m etals in the tab l e for m stable complexes with either m e thyl orethyl groups. In addi tion to m e thyl/ethyl
com pounds, m etals can also be in corporated b i ochemi cally in to stab le org anom etallic
compounds (e.g., As into lipids and Se into amino acids). Organom etallic environ m ental trans f orm ations m ay af f

ect both the m obility and toxicity of m etals. The transform ati on rates and the organom e tall ic products are dependent on environm ental conditions and the population of mi croorganism s available. For example,
m ethylation/dem ethylation rates are dependent on the speciation of the metal, the mi crobia l community, the environm ental variables (e.g., pH , temp erature, reduction oxidation potential, organic m a tter, dissolved oxygen, nutrient availability, salinity, complexing agents), and the distribution of the m etal betw een co mpartm ents (sedim ent, water, gas eo u s). The in terrelatedness of these processes has made research into unravel ing the factors controlling net
m e thylation difficult and, to date , incomplete. However, some ge neral trends can be predicted with som e certainty, a s discussed in this se ction. 3-4

### Table 3-3. Metals/metalloids involve d in methylation processes
Process Metals a f fected Environm entally stable organom etallics Si, Ge, Sn, Pb, Hg, As, Sb, Se Abiotic chemical m e thylation Hg, Pb, Sn Abiotic dem e thylation Sn, Pb Biotic m e thylation As, Cd, Hg, Pb, Se, Sn Biotic dem e thylation As, Hg, Sn, Pb Source: Bodek et al. (1988).

3.1.2.1. Biotic Methylation Transformations Biotic m e thylation occurs when organism s, prim arily mi croorganism s, transfer alkyl groups to bioavailable m e tals. In general, it is thought that an aerobic sulfate-re ducing bacteria are the principal m e thylators in freshwater and estuarine environm ents. However, m e thylation rates are no t always correlated with s ulfate-reduc ing bacte r ia. Not all su lf ate- reduc ing bacte r ia are capable of m e thylating, and the efficiency of m e thylation is dependent on the activity and structure of the bacterial community. Other bact eria m a y be involved in m e thylation. Biotic
m ethylation occurs predominantly inthe sed i m ent column ; however, because the water column by volum e is mu ch larger, water column m e thylation is important. Maximu m m ethylation rates typically o ccur at the redox boundary, which varies seasonally and frequently coincides with the sedi m ent-water interface (U llrich et al., 2001). Methylationrates d ecr ea se with in cre asing sed i me nt depth, probably due to a decrease in biotic habitat. Microorganism s m a y also dem e thylat e (or dealkylate) organom etallic compounds. Microbial-mediated transfor m a tions are fre quently the m o st important environm ental organom etallic p r ocess es . Generally, as the am ountof organic m a terial in creases in a s y stem the
mi crobial populations also increase. High temperatures and anaerobic conditions generally favor m e tal-m e thylation for m ation, and dem e thylation processes are gene rally favored under low tem p eratures and/or aerobic conditions. S tudies on the effects of pH are not consiste nt. Interconnected param e ters include pH e f f

ects on the mi crobia l c o mmunities a nd ef f

ects on the spec iation distribu tion of the
m etals/m etalloids in the water and th e sedim ent as well as adsorption rates. Organom e t allic
compounds appear to increase in th e water column in low pH envi ronments, but this m a y be due to re lease of m ethylated m etals f rom the so il and subsequent depletion of organom etallic
compounds in the soil. Therefor e, pH effects on net m e thylation in a system are not fully understood. In freshwater ecosystem s, where sulf ate concentrations are typically low, increase
3-5

### in sulfate co ncentration increas es m e thylation rates. However, in reducing environm ents, increasing sulfide concentration decreases m e thylation rates. General trends in
m ethylation/dem ethylation rate s are outlin ed in Table 3-4.
Table 3-4. General trends of enviro nmental factors affecting rates of methyla t ion/demethy l ation Organometa llic transform a tions Temp erature pH SO 4 -2 Organic matter Redox Salinity High Low High Low High Oxic Anoxic High Net m e thy lation ↑ ↓ ? ? ? ? ↓ ↑ ↓ Methy lati on a q ↑ ↓ ↓ ? ↑ ? ↓ ↓↑ ↓ ↑ ↓ Methy l ati on s e d ↑ ? ↑ ↓ ↓ ? ? ? ? Demethy lation ↓ ↑ ↓ ↑ ? ? ↑ ↓ ? 

### ↑ indicates an increase in rate. ↓ indicates a decrease in rate. ? indicates conflicting data or insuffici ent data to indicate a likely trend.

The inhibitory effect of sulfide is probably not due to m e tal sulfide form ation but, rather, to the f o rm ation of less bioavailable m etal-su lf ur comp lexes. High organic m a tter m a y increase abiotic m ethylation thro ugh humi c/fulvic m etal reactions; ho wever, this m echanis m is poorly understood and confounded because bi otic m e thylation rates m a y in crease in environ m ents with high organic m a tter. In ecosystem s with hi gh dissolved organic carbon (DOC) concentrations, DOC may bind with m e tals/m etallo ids, renderi n g them unavailable and thereby reducing biotic
m ethylation rates.

3.1.2.2. Abiotic Transformations Som e organom e tallic compounds, those withelectronegativities greater than 1.7, are stable under environm ental conditions. In a few instances, m e tals can undergo abiotic
m e thylation processes. Carbon-m e tal bonds with m o re polar (m etal electronegatives <1.7) will undergo hydrolysis (reaction with water). Abiotic chemical m e thylation can occur by three
m echanis m s : transm ethylation reactions betwee n m e rcury and tin/lead al kyls, reactions with humi c/fulvic substances, and photochemi cal reactions.

3-6

### 3.1.3. Aquatic Chemistry Exposure and risks associated with metals in aquatic environm ents depend on the form s of the m e tals and on the factors th at influence the presence of these form s as well as o n the fate and transpo rt of the m etals.

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-s ized particles; biof ixation occurs when biological processes (usually involving mi croorganism s or plants) resu lt in the b i nding of m etals to so lid m aterials; interactions with natural organic m atter can also result in sorp tion. In addition to these factors, sorption is influenced by changes in pH, oxi dation potential, salinit y, concentrations of
competing ions, the nature of sorbent phases and th eir surface areas, and s u rf ace site d e nsities. n Factors Affe cting Sorpti on of Metals Speciatio n /co m plexation Preci p itation Colloid formation Biofixation Interactio ns with natural o r gani c matter 

### on Cha nge s in p H Oxidation pot ential Salinity 

### f Comp eting io ns Nature of so rbent pha se s Surface site d ensitie s



### 3.1.3.2. Speciation and Complexes Meta l spec ie s dissolv ed in water m ay occur as free ions, or aquo-ions , or as com plexes. Free m e tal cations are generally surrounded by c oordinating w a ter m o lecules and so have been term ed “aquocation s,” although by convention th e water m o lecules are ig nored when writing chemical reactions involving m etal cations. The total an alytical concen tration of a given m etal in wate r is the sum of the concentrations of its free ion and its complexes an d any m etal associated w ith suspend e d solids, whether organic or mi neral. For example, the to tal m o lal concentr ation of lead, Σ Pb, in a natural water mi ght equal:

Σ

Pb = mPb 2+ 

### + mPbOH + 

### + mPbCO 3 o 

### + mPbHCO 3 + 

### + mPbSO 4 o 

### + mPb(suspended solids)

In m ost natural waters, the concentration of free lead ion, mP b 2+ 

### , is less than the sum of the concentrations of its complexes, which inthis case are lead complexes with hydroxyl, carbonate, bicarbonate, and sulfate io ns. Other metals that are found in natural waters most often
3-7

### as complexes and not as free ions include Al 3+ 

### , Ag + 

### , Cu 2+ 

### , Fe 3+ 

### , and Hg 2+



### . The m etalloids As and Se and the m e tals Cr, Mo, Sb, and V occu r most often in aerobic waters and soils as covalently bonded to oxygen in oxyanions. U nder oxidizing conditions, th ese include arsenate, selenate, chrom a te, m o lybdate, and vanadate, which them selves are comp l e xes. Com p lexes that inco rporate m etals play a m a jor role in con tro lling the availab ility and fate of m etals in the environm en t. Increasing the fraction of a m e tal thatis com p lexed incre ases the solubility of mi nerals of that m e tal (Langm uir, 1997). For example, the solubility of lead sulfate is related to the molal c oncentrations of free lead and su lfate ions through the expression below:

K sp 

### = ( γPb 

### .mPb 2+ 

### )( γ SO4 

### .mSO 4 2- 

### )

where the term s γ Pb 

### and γ SO4 

### are the activity coefficients of the ions. The product of the ion activ ity co efficient and the m o lal concentration of each species equals th e activity of the ion. This equation shows that for a given total Pb c oncentration, the gr eater the am ount of Pb that is
complexed, the lower the concentration of free Pb ion. This m eans that as the extent of Pb
com p lexing incre ases, th e tota l Pb co ncentr ation must also in creas e to r ea ch satur atio n equilibrium with lead sulfate. In other words, m etal comp lexing incre ases tota l m etal solubility. Metal complexing also has a direct influen ce on m e tal adsorption to organic m a tter or
mi neral surfaces. For ex ample, m etal carbon ate, sulfate, and fluoride co mplexes are usually poorly adsorbed, whereas m e tal hydroxide comple xes are strongly adsorbed (Langmuir, 1997). In summ ary, m etal com p lexing gen er ally incre ase s the solub i lity and m obility of m etals in surface and ground waters (with the excep tion of m etal hydro x ide com p lexes). For m a ny metals, the free m e tal ion is thought to be the prim ar y m e tal species that causes toxicity to aquatic organism s. This is consiste nt with the free ion activity model (FIAM), which carries the assump tion that the free or aquo-ion is the m ost biologically active form of the dissolved m e tal. If this is true, the key param e ters that can modify the degree of toxicity are those that af fect speciation, su ch as pH and the amount of i norganic and organic ligands (e.g., DOC) that can for m m e tal comp lexes and so provid e alternative binding s ites for the m etal ion. Metal toxicity also will be a ffected by other dissolved ions (e.g., Na, Ca) that comp ete with
m e tals for binding sites o n the gills of fish or on th e respiratory surfaces of other aqu a tic organism s. However, it should be noted that m a ny m e tal complexes (whether bound to organic or inorganic ligands) are reversib le, particularly if environm enta l conditions change (e.g., pH decreases).

3-8

### 3.1.3.3. Importance of pH and Redox Conditions The pH is probably the single m o st important variable that influen ces the behavior of
m e tals in the environm ent. Thus, m e tal complexes with sulfate, fluoride, chloride, and phosphate are m ost stable and important below pH 7, whereas m e tal carbonate and hydroxide
comp lexes becom e increasingly m o re important above pH 6–8. Also, as discussed in Section 3.1.6, hydroge n ion comp etes with m e tal cations for adsorption sites, so that adsorption of m e tal cations by hydrous ferric oxide (HFO), for example, is low in a c id system s but inc r ease s with inc r eas ing pH. In contrast, oxyanions of As, Mo, Se, and Cr tend to be desorb ed f r om HFO with incre a sing pH be cause of co mpetition b e tween the oxyanions and OH- ion for sorption sites. Furtherm ore, the solubility of most m e tal-containing
mi nerals is greatest under acid conditions, decrea sing with increasing pH. Figure 3-1 shows the locus of m e a s ured values of oxidation potential (E h) and pH in aquatic system s. The pr incipal controls on Eh are atm ospheric oxygen and orga nic m a tter. High Eh ( oxidizing or aerobic) conditions are m a intained in the atmosphere and in m ost surface waters and shallow soils in contact withatm o spheric oxygen. The lowest E hvalues andreducing o r anaerob ic co nditions are found in water-logged soils a nd sedim e nts that contain organi c m atter and in ground waters that con t ain a few millig ram s per liter or m o re of DOC. Interm ediate Eh conditions are found in waters and s e dim e nts that are on ly partially ox idi zed becaus e of their relative iso l ationfrom the atm o sphere. Measured E hvalues m ay not agree with Eh valu es com puted from the concentrations of redox-sensitiv e species. The difference between m e asured and computed Eh values is discussed at length by Stu mm and Morgan (1996) and Langmuir (1997).

3.1.4. Ground Water and Metals Mobility Site-specific nonorganic m e tal health risk as sessm e nts for EPA programs often are used to predict the rate of m o vem e nt of metals through soils and their subsequent m ove m e nt and concentrations in ground water. The prim ary processes governing the environm ental fate and transpo rt of m e tals in the subsurface are advec tio n, dispersion , m a trix diffusion, and retard ation (U.S. EPA, 1994b). Advection and dispersion are functions of the system rather than of the contaminant. Matrix dif f usion, which is a functi on of the contaminant, is relatively unimportant and is omitted in m ost model transport algorithm s. Retardation depend s on a num be r of factors (Langmuir, 1997; U.S. EPA, 1994b) and m a y i nvolve or be affected by the following:

• Sorption - The attachment of m e tal species to mineral surfaces or other s urfaces.

• Speciation - The distribution of a given constituent am ong its possible chemical for m s, including m e tal comp lexes, which have differing tendencies to be adsorbed or desorbed.

3-9

### Figure 3-1. Approximate position s of so me n a tural environments in terms o f Eh and pH. The dashedline repre se n ts the limits of m easurem ents in natural environm ents. Source: Baas-Becking et al. (1960).

• Precipitatio n - The proc ess by whic h dissolved species exce ed the so lubility limits of their solids, so that som e of the species precipitate from solution. Wh en a m etal species reaches mi neral saturation, addition of further am ounts of the species to solution are precip itated, not adsorbed.

• Colloid for m ation - The process of for ming colloids and th e associ ation of m etal species with them

. The m e tals m a y be sorb ed or coprecipitated with colloidal-siz ed particle s.

• Biofixation - The binding of m e tals to solid m a te rials due to the interactions of mi croorganism s or plants.

• Natural org a nic matter- (NOM) in tera ction s.

3-10

### • Other processes - Oth e r proces ses include chan ge s in pH, oxidation potential, salin ity, con centr ation s of comp eting ions , the nature of sorb ent phases and their surface areas, and surface site densities.
3.1.4.1. Applicat ion of Partit ion Coefficient s

t

o

Metal Mo bility in Gr ound Water Owing to the com p lexity and mu ltiplic ity of the processe s in volved, re co urse is of ten
m a de to the use of a single partition or distribu tio n coef f i cient that de scri b e s the deg r e e to which the contaminant’s trans port is re tard ed relativ e to water. This approach starts with defining the retardation factor:



### ν R f ν p = c where: R f 

### = the retardation facto r , ν p 

### i= the velocity of water through a control volum e, and ν c 

### = the velocity of contaminant through a control volum e. Langmuir (1997) noted that the re tardation factor is related to the distribution coefficient through the below expression:



## R f 

## = 1 + ρ b 

## K d 

## n e


### where: ρ b 

### is the porous m e dia bulk density, and n e

### is the effective porosity at saturation given as a volum e fra c tion.
Laboratory adsorption studies of ten find that, in simple system s, the value of log K d

### for
m etal adsor p tion in crea ses lin early with pH. For exam ple, for Zn 2+ 

### adsorption by H F O, Langmuir (1997) noted that adsorp tion followed the log equation:
log K d 

### = ! 5.48 + 1.77 pH
The properties and applications of partition or distribution coeffici ents f o r m etals are
comprehensively review ed in EPA (1999a). Th ese coefficients m a y be obtained from the literature, es tim ated usin g m athem atical m ode ls, or m easured. Properly designed column

3-11

### experim e nts generally give results m ore representative of the in situ behavior of soils and sedim e nts than do batch tests (Langmuir, 1997). Models using partition distribution coefficients (K d 

### ) have sig n ificant inaccuracies for
m e tals, and the application of si ngle partition coefficient values for individual m e tals should be limited to site-specific assessm ents (when they can be calibrated f o r a spe cif ic site) ; to regiona land national-scale studies where bounds of potential K d

### values are adequate; or reasonably repres entative single values that hav e sufficien t functionality built in to account for variab ility 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.go v/ceampubl/mmedia/ minteq/index.htm or http://www.lwr.kth.s e/english/OurSoftware/Vminteq/).

For screen in g assessm ents:

• Partition coefficients have been tabulate d 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 hav e large v ariability and sele cting m eaningf ul values is dif f i cult), m ercuric chloride (58,000), and Hg (1,000) (U .S. EPA, 1999b).

• In sim p le system s, the value of log K d

### for m e tal cation adsorption usually increases linearly with pH, whereas the value of log K d

### generally decreases with pH for anion adsorption.
For definitiv e assessm ents:

• It m ay be possible to es tim ate m etal adso rption with som e accuracy witho u t having to m easure it in the field, depending on the in form ation availab l e on a spec ific soil, surface-water, or ground -water system. To do so, at mi ni mu m, the risk assessor will need to determine the amounts and surface areas of the potentially so rbing m a terials (e.g., m e tal oxides, clays, and organic m a tter [OM]) in a soil or sedim e nt or in suspension in a stream , and the detailed chemical composition of the water, especially its pH and m e tal concentration. Literature inform ation can then be used to estim a te the sorption properties of these m a terials fo r use in the diffuse layer (DL) sorption model.

• If greater accuracy o rsite specificity is required, it m a y be necessary to m e asure m etal adsorp tion in labo ratory experi m ents designed to param e terize the DL m odel for application to a specific study area. The risk assessor could run experim e nts (batch tes t s) that attemp t to reprodu ce th e composition of waters and sorbing solids in the study area.

3-12

### 3.1.5. Sediment Chemistry In addition to the challenges posed by m e tal chemistry, the sedim e ntary environm ent is comp lex and often highly heterogeneous. Fortunately, generalizations can be m a de about the sedim e ntary environm ent in order to progress toward a
m e thod for r i sk assessm e nt. This brief review summ arizes inform ation on the
composition of sedim e nt s; processes that act on sedim ents and their m etal burden; and the chemistry of the sed i m entary env i ron m ent that inf l uenc es the f a te, bioavailability, and effects of metals. It is importa nt to consider these factors inlight of the aim of estim a ting potential b i o l ogical effects of m e tals in sedim e nts. Sedim ent solids c an holdup to a million tim es m o re m etal than an equiva lent volum e of water. The exact proportion of a chemical held by sedim e nt relative to water is a function of a
m e tal’s chemistry as well as the chemistr y of the sedim e nt solid and the surrounding environm ent. Further, th is distribu tio n is dynami c (Diam ond a nd Mudroch , 1990). Because of their large capacity to “h old” m e tals, sedim ents have been charac terized as “sinks.” S e dim e nts
can serve as temporary sinks from which som e of the m e tal can enter ecological and h u m a n food webs through several routes (e.g., Diam ond, 1995), prim arily through accumulation in benthic organism s. Thes e organism s include thos e that f u lly o r partia lly live in the sed i m e nts (e.g., tubificids, chironomi ds, tricho pteran larvae) or those that f eed from the sediment bed (e.g., suckers, carp). Som e or ganism s obtain their ch emical dose from both pela gic and benthic routes (e.g., lake whitefish, walleye), but because of highchemical concentrations in sedim ents, the benthic route can be the do mi nant route of uptake (M orrison et al., 2000). For hum ans, the route of entry of m e tals from the sedim e nts is through water used for drinking, bathing, and swimmi ng. The availability of these metals is m e diated by sedim e ntwater exch ange process es thatcan r esult in the r eleas e or r e mobilizationof chemicals f r om the sedim e nt bed. However, owing to the ability of Hg to bioaccumulate in its m onomethyl form , fish consumption is the critical rout e of exposure to this m e tal for hum ans . 2 

### Many important chemical reactions involving th e m e tals of concern occur in the finegrained m aterials th at accumu late in the deep parts of water bodies. The controlling factors or
m aster variables that influence m etal 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 μ m) to silt (2–63 μ m) to clay (<2 μ m). Because clays have more active binding sites than do the other grain sizes and because of their highsurface area-to-volume ratio, clays –or finegrained particles--are of greatest significance in terms of metal binding.

2 M ari ne bi ot a c a n al s o be a si g n i f i cant ro ut e o f e x p o s u ret o ar seni c i n i t s or gani c f o rm s, su ch as m e th y l ar soni c aci d; howe ver, thesearse nic s p ecies are significantly le ss t oxi c th an th ei n or gani c f o rm s (Fo w l e r, 1 9 8 3 )

. 

### 3-13

### the sed i m e nts will reveal decreasing sedim e nt porosity and concentration s of dissolved oxygen because oxy gen is consu m ed as organic m atter d eco mposes. The pH is often relatively constan t
or m a y decrease with depth, but alkalinity m a y increase owing to mi neralization of organic
m a tter (Stumm and Morgan, 1996). As dissolvedoxygen is consum ed, a naerobic mi crobes use other electro n accepto rs in redox or o x idation-redu ction reactions in the o r der of nitrate, ferric iron, amm o nium , sulfate, and bicarbonate to produce carbon dioxide, ammonia, sulfide, and
m e thane. The risk assessor can ascertain the redox st atus of the sediments by m e asuring the concentration of dissolved oxygen or other redox-se nsitive species, by calculating pE (a m easure of

electron a v ailab i lity insolution ) , o r by m easur ing Eh (millivolts) o rth e elec trom otive f o rce of the pore water solution. The zone of transition from oxic to anoxic conditions is the redoxycline, which can migrate vertically, depending on th e mi xing of the overlying water column (e.g., Diam ond and Mudroch, 1990). For exam ple, th e redoxycline m a y be 5-10 cm below the sedim ent-water interface in a well -ox ygenated o l igotrophic lake or riv e r, but it m a y be above the sedim ent-water interface in a therm ally stratified eutrophic lak e or riv e r.
3.1.5.1. Metal Chemistry in Sediments In this discussion, the risk a ssessor should consider two pools of m e tals. The first pool consists of m e tals that exist as aqueous (o r dissolved) species bound to colloids or DOM and those bound to sedim e nt particles through an exchangeable binding process. This pool is often referred to as the “exchangeable” or “labile” poo l. The second pool consists of m e tals found within the mineral m a trix of the sedim e nt solids. This pool is largely unavailable to biota, and its release will occur over g e ologic tim e scales, thro ugh diagenetic p r ocesses. Because th e latter pool is largely unavailable, only th e exchangeable pool of m e tals 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 proc esses as well as m etals re lease dinto th e environm ent as a result of anthropogenic activities. The exchangeable pool of m e tals is subject to speci ation in the aqueous phase (e.g., within th e p o re wate r) and sorption to solid phas es, where sorption is a general term t hat includes adsorption (the accumulation of m a tter at the soli d

- water interface or a two-dim ensional proces s) and absorption (inclusion in a three-dim e ns ional m a trix) (Stumm a nd Morgan, 1996). Here, speciation re f ers to the d i stribu ti on of m etal spec ies in a particular sample or m a trix or species distribution (Templeton et al., 2000 ). In the aqu e ous phase, m etal will react or b i nd with dissolved lig ands acco rding to the pH , Eh, ionic strength, and abundance o f ligands (see Section 3.1.3, Aquatic Chemistry). The concentration of m e tal in the dissolved phase is controlled by sorption to the solid phase. Although sorption can occur rapidly, desorp tion or dissolution of m e tal from t h e solid phase m ay be a two-phas e process, w h ere th e second phase is rate limitin g (e.g., Nyffeler, 1986 ; 3-14

### Santschi et al., 1986). If the kine tic limitation of reactions are not consid ered, the distribution of
m e tals am ong aqueous species and between th e aqueous phase and the solid phase can be estim ated by using one of the speciation m odels

. Several speciation/comp lexation m odels are available to perform this cal culation, such as MINEQL+ (Sc h echer and McAvoy, 2001), the W i nder m ere Humi c Aqueous Model (WHAM) (Lofts and Tipping, 1998), and MINTEQA2 (Allison et al., 1991). T h ese m odels work well u nder oxic co nditions, but estim ates of m etal binding are less reliable under anoxi c conditions, where m e tal concentrations are m ost often contro lled b y the solub i lity of m etal sulf ides. Fu rtherm ore, in som e circum stances equilib rium
m a y not be achieved, particularly w hen the redo xycline m oves m o re quickly than the rate of
m e tal reaction or when the reaction is governed by mi crobial processes, as occurs with the
m ethylation of Hg or As. In oxic sed i m ent pore waters (above the re doxy cline), m etals will ex ist as aqueous species, that is, as freely disso lved ions or m etal com p lexes (e .g., phosphate, sulfate, or carbonate
complexes), and associated with colloids. Solid-phase reactions are controlled by iron oxyhydroxides and m a nganese oxides that m a y exist as colloids, sedim e nt particles or surface coating s of particle s, OM that m ay also ex ist a s collo ids or c o at sed i m e nt particles, and clay collo ids and particle s. As Eh declines, the solid-phase m a nganese oxi des are the first to be reduced and thereby dissolve, wh ich releas es m etals that h ave been sorbed or coprecipitated with these mi nerals. Som e of the m etals relea sed into the por e water m a y 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 consum ed a nd the S is converted to sulfide, m e tal concentrations in pore w a ters again drop as soli d- phase m etal- sulf ides ar e for m ed (see discussion below about the role of aci d-volatile sulfides [AVSs] in regulating toxicity). As a result of redox chemistry, m e tals can undergo seasonal redox-driven cycling between the water column and sedim e nts or with in the sedim e nts, depending on the position of the redoxycline. The stages in th e cycling are, first, the adsorp tion or coprecipitation of m e tals with iron and m a nganese hydroxides under oxidizing conditions; then with the developm ent of moderately reducing conditions, the reduction and dissolution of th e m anganese and iron oxyhydroxides and consequent release of the associated m e tals into the water or pore water; followed by their diffusion upward toward the zone of low m e tal concentrations under oxidizing conditions. It is also possible for dissolved m e tals to diffuse downward toward the zone of low
m e tal concentration owing to their precipitation as sulfides. As a re sult of this vertical cycling, the depth profile of m e tals in pore water m a y not m a tch 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. Unde r these conditions, the solid-ph ase sedim ent prof ile ref l ec ts this r eworking rather than the historical record of m e tal loadings (MacDonald et al., 2000). 3-15

### The pH controls m e tal speciation and bindi ng by affecting the species distribution of dissolved ligands (e.g., phosphate, sulfate, carbonate, humic subs tances) and the surface charge of binding sites on DOM and solid phases such as iron oxyhydroxides. G e nerally, at low pH, when surface sites are protonated, th e sorption of cation i c m e tals d ecreases, and, hence, m e tal mobility increases. Th e converse o c curs at h i gh pH, which results in low m etal solub ility and greater sorption. The patterns of dissolution and sorption are revers ed for m e talloids, such as As, that ex ist as anionic species.

3.1.5.2. Estimating Me tal Distribu tion in Sediments A m ain objective in te rms of assessin g the hazard or risk posed by m e tals in sedim e nts is estim ating the am ount of m e tal that is potentially bioavailable. The bioavailability of m e tals in sedim e nts is a function of their distribution between the dissolved a nd solid phases, with dissolved m e tals in porewater generally cons idered to be the m ost bioavailable fraction. Accordingly, several m e thods have been develo ped to estim ate the distribution of m e t a ls am ong dissolved and solid phases in sedim e nts. Th ese m e thods have been thoroughly reviewed by Mudroch et al. (1999, 1997). A lthough bioavailability is also a function of aqueous phase speciation, limited research has been conducted to estim ate metal speciation in pore w a ters. Generally, for the purpose of ecological risk asse ssm ents, the exposure of benthic organism s to sedim e nt-associated m e tal is assum e d to be proportio n al to the m etal concentr ation s in inte rstitial water, although som e st udies indicate that uptake from overlying water (Hare et al., 2003; Roy and Hare, 1999) or ingested sedim e nt m a y be a significant source of body burdens of m e tals. Distribution of m e tals in sedim e nt por e waters m a y be determi ned by field
m easure m ents, experim ental m ethods, and m athem a tical m o deling, with the latter also requiring
som e field m easure m ents. Concentrations of m e tals in pore waters m ay be dete rmined in the field by using pore water dialysis chambers or pe epers and by m e thods that separate the solid phase from t he pore water, although the latter have been shown to be less reliable (Mudroch et al., 1997). S e veral extraction schemes have beendeveloped to determine the distribution of
m e tal a m ong operationally defined f r actions (e .g., Forstner, 1995; Te ssier et al., 1979). However, sequential extraction m e thods do not cl eanly distinguish the oc currence and speciation of different form s of m e tals in sedim e nts and soils (Tye et al., 2003; Tack and Verloo, 1999). Other experim ental m ethods includ e leaching test s (e.g., Reuther, 1999). The results of any of these m e thods are concentrations of m e tals in pore water, which can be related to toxicity benchm arks. Because of the need to d e velop Sedim e nt Quality Guideline s (SQGs) for m e tals that explicitly address toxicity and are based on read ily m easured param e ters, several m e thods have been developed. For oxic sedim e nts, Tessier et al. (1993, 1989, 1984; Tessier, 1992) compiled partition coe f f i cients of m etals th at were derived from field studies of freshwater sedim e nts. The
3-16

### partition coefficients are dependent o n pH (becau se Eh is heldconstan t ) an d are generally linear over a range of pore water pH values (see above discussion under gr ound-water chemistry, Section 3.1.4 for the theoretical basis for de velopm ent of partition coefficients). Speciation/complexation m odels al so m a y be used to estim a te fractions of dissolved and bound
m etal species. These m o dels rely on m easure m ents of pH, dissolved oxygen, or Eh to establish redox conditions. The models assum e that solid -phase binding is governed by sorption to iron and m a nganese oxides. Model estim ates are less re liable when other solid -phase substrates are domi nant (e.g., clay mi nerals) and are a function of the av ailability and accur acy of the stability constants for the m e tal-ligand reactions that are used in th e c alculation s. Model es tim ates ar e less reliable when other solid-pha se substrates are domi nant (e.g., clay mi nerals), and they are a function of the availability and accu racy of the stability cons tant s for the m etal-ligandreactions that are usedin the calcu l ations. For anoxic s e dim e nts, the availability of sulf ide contro ls m etal dis tributio n and solubility. Operationally, AVSs ( m ainly iron m onosulfide) have been consider ed as a m easure of reactive sulfides (Forstner, 1995). Studi es have dem onstrated an invers e relationship between sedim e nt acute toxicity and AVSs for m arine and freshwater sedim e nt s (Di Toro et al., 1992, 1990; Ankley et al., 1991) as well as between pore wate r concentrations and A V Ss (Brum b augh et al., 1994; Casas and Creceliu s, 1994). As a screeni ng-level tool, th e acute toxicity of anoxic sedim e nts can be assessed by determining the ra tio of AVSs to simu ltan eously ex tra cted m etal (SEM). Low sedim e nt toxicity is indicatedwhen AVSs are in excess (AVS > SEM), which im plies suf f i cient capac ity of the AVS to bind ess e ntia lly a ll f r ee m etal. This topic is f u rther discussed in Section 5.

3.1.6. Soil Chemistry The cation i c m etals occu r natur ally in soils a s oxides and hydroxides (Fe, Mn, Al); to a lesser extent they can occur as carbonates, phosph ates, and sulfates; and inreducing (usually wet or waterlogged) soils as sulfides, which are highl y insoluble. The soil par am eters importan t in affecting sorption and p r ecipitation reactions an dthe extent o f their influence—and th us contaminant bioavailability—depend on the intrinsic properties of the contaminants. In the soil environm ent, m etals can exist as cations (having a positive ch ar ge), anions (having a negative charge), or neutral species (havi ng a zero charge). Their form significantly affects their sorption, solubility, and m obility. For exam ple, m o st so ils are ch ief l y negativ ely c h arged; thus , m etal cations have a higher propensity to be sorbed by soil particles than do m e tal anions (U.S. EPA, 2003e). Cationic m etals can react withino r ganic soil cons titu ents (e.g. , carbonates, sulfates, hydroxides, sulfides), forming aqueous complexes, which m a y be adsorbed or precipitated in
mi neral form. Most complexation and precipit ation reactions are pH dependent (U.S. EPA, 2003e). 3-17

### As, Cr, Se, and V complex with O and typica lly exist as anio nic spec ies u nder m o st environm entally relevant scenarios (Bohn et al., 1985; Lindsay, 1979). The m ost common for m s of As are arsenate (As(V)) and ar sen i te (arsen ic( I II) ), which ar e pres ent insoil so lutio n in the for m of AsO 4 3- 

### and AsO 3 3- 

### , respectively. The chemistry of arsenic resem bles that of phosphate (Barber, 1995; Bohn et al., 1985). Cr can exist as chrom a te (Cr(VI) or CrO 4 2- 

### ), which is usually considered more soluble, m obile, and bioavailab le than the sparing l y so l uble chromite (Cr(III)), which is norm a lly prese nt in soil as the prec ipitate Cr(OH) 3 

### (Barnhart, 1997; Jam e s et al., 1997). Similarly, Se can be present as selenates (SeO 4 2- 

### ) and selen ite s (SeO 3 2- 

### ). For V, vanadate (VO 4 3- 

### ) is the m ost comm on form (U.S. EPA, 2003e). Metals can exist in the pore water as aquo- ions or soluble complexes. The bonding of
m etal specie s to soil particle s can r an g e f r om ionic to cova lent. For m o st soils in the U n ited States, negatively charged sites are m o re plen tif ul; less than 5 % of the total availab l e charge on the soil surf ace is pos itively charged. Metals existing as cationic sp ecies have a greater propensity to associate with such soils. This m ake s them less bioavailable, but it also results in greater lo ading of m e tals into th e soil ecosyst em . Anionic m e tals generally m ove into pore water—and so are m ore bioavailable—but leach out of the system mu chmore rapidly. In summary, soil pH and av ailab i lity of charged s ites on soil su rfaces are th e prim ary soil factors controlling release of m e tals to pore water and, subsequently, bi oavailability (U.S. EPA, 2003e).

3.1.6.1. Ke y Paramete rs Affec t ing Met al Bioavailability in Soils From the preceding overview of how the m e ta ls and m etal com pounds interact with soil constituents, it is cle ar th at soil play s a ve ry sign if icant role inreducing the potential bioavailability of m e tals in the environm ent. Given the typ es of contaminant-so il inte r action s presented, th e prim ary soil facto r s controlling the potential bioavail ability of m etals ar e soil pH, the availab i lity and ch aracter of sorp tion sites on soil su rfaces , the content of Iron and aluminum oxyhydroxides and soil organic m a tter, and least important, the soil clay mi neral content. The following discussion briefly details the key soil p aram eters af f

ecting the v arious con t aminants availability to the pore w a ter. Soil pH is often term ed the m aster soil variab le b ecause it co ntrols v i rtually all asp ects of contaminant and biological processes in soil

. These processes in clude dissolution and precip itationof m etal solid phases, c o mp lexationand acid

- ba se rea ction s of m etal spe cies, and
m etal sorption as well a s mi crobial activ ity. Inc r eas ing soil pH also resu lts in an inc r ease in the
num ber of negatively charged soil sites, with a concomitant decrease inthe positively charged sites. Th ere f ore, incre as i ng the so il pH incr eases the sorption and rem o val from pore water (Bohn et al., 1985).

3-18

### 3.1.6.2. Adsorption Behavior of th e Metals of Concern 3.1.6.2.1. Surface area and surface density. In porous m e dia, the m ost important sorbent solids for m e tals are oxyhydroxides of iron and m a ngane se. Their important surface properties are discussed in Langmuir et al. (2003)

. For a given weight of sorb ent, m etal sorption capacity is proportional to surface area and su rface site dens ity. The greatest surface site d e nsities (positively or negatively char ged sites) are those of organi c material and the oxyhydroxides. These phases are the strongest and m ost important so rbents of trace m e tals. Except for kaolinite, the clays (0. 02-2 mmol sites / g) hav e a surface ch ar ge th at is largely indep e ndent of pH, whereas the surface charge of org anic m atter and the oxy hydroxides is strong ly p H dependent. 3.1.6.2.2. Organic matter (organic carbon) con tent. Organic m a tter includes plant and anim al
rem a ins in various stages of decomp osition, cells and tissues of soil organism s, and substances exuded from plant roots and soil mi crobes (S umner, 2000). O r ganic m atter is prim arily
composed of carbon, oxygen, and minor am ounts of nitrogen (N) and phosphorus (P). On average, approxim a tely 58% of organic m a tter is organic carbon. Organi c m a tter in soils ranges from <1% for a sandy soil to alm o st 100% for a p eat soil, with m o st soils having organic m a tter contents <1 0% (Bohn et al., 1985). Also, organic m atter con t ent is usuall y higher in surface soils or in the root zone and decreases with depth in th e soil p r of ile

. The organic acid functional groups typically present in organic matter have a high affinity for m e tal cations.

3.1.6.3. Aging of Metals in Soil A distinction should be m a de betwe e n persiste nce of total m e tals insoil and persistence of bioavailable form s of the m etal. As m etals age in soils, they decrease in bioavailability. The aging process is partially revers ible if environm ental param eters change (e.g., pH decreases), although a p o rtion of the m e tal ions will be s ecure ly entrapped in the so il particle lattice and not available to be resolub i lized. It h a s b een we ll doc um ented thatm etal che mistry in solutions freshly added or spiked into soils vary from m e tal form s in fi eld-con t aminated so ils. Typically, the m etal contaminant pool requires tim e to diffuse into mi cro- or nanopores and to be absorbed into organic m a tter and soil particles. These sl ow reactions are attribut ed to mi cropore diffusion, occlusion in solid phases by (c o)precipitation and (co)flocculation, and cavity entrapm e nt. Although the slow reactions play a key role in me tal bioavailability, their rates, m e chanism s, and controlling f actors have not been comprehensively elucidated. Ev idence of aging processes is provided by studies of metal extract ability and lability (Tye et al., 2003; Ha m on et al., 1998). It has been frequently observed that easily extrac table pools revert with tim e (~1 year) to more strongly bound for m s. Isotopic dilution provides a useful way to quantify changes associated with progressive attenuati on of m e tals in soil. After 1 year ag ing reaction s are alm o st com p lete and are m ostly reversible. At present, inform ation regarding the aging reactions of dif f

erent

3-19

### m etals and m etalloids and sorbing s o lids, is very limited, so it is no t pos sibl e to generalize which
m etal(s) age the f astes t o rwith gre ate r/less rev ers ibility.

3.1.6.4. Dissolution an d Transformation of Metals The dissolution and transfor m a ti on of a m e tal compound in soil is related to a series of chemical and physical properties characteristic of the com p ound itself and of the soil. Environm ental param e ters such as temperature a nd humi dity have a strong influence on the rate of transformation. When assessing the transform a tion of a compound in so il, the risk assessor should keep in mi nd that aging reactions m a y take place at th e sam e tim e as tran sformation and dissolution. Wh en m etal salts are ad ded to soil, the for m of t he salt dictates the rate and am ount of soluble metal that will for m in the pore water. Insoluble form s of m etals (e.g., vanadium pentoxide [ V 2 

### O 5 

### ] ) will transf orm to soluble f r e e ion (V) at a slower rate th an will so luble m etal salts (e.g., N a 3 

### VO 4 

### ). However, the rate of form a tion of the free ion is not proportional to the dissolu tion rate of the sa lt becaus e ag ing rea ction s will rem o ve the free io n from the pore wate r. The risk as s essor shou ldconsider the relative rates of dissolution and aging simultaneously to accurately p r edict pore water concentrations.

3.1.6.5. Soil Metal Tra n

sfer to Pla n

ts The “soil-plant barrier” concept was intr oduced to communicate how m e tal addition, soil chemistry, and plant chemistry affect risk to an im als from m e t a ls mi xed in soil (Chaney, 1980). Reactions and processes that take place at the so il-p l ant b arrie r are influenced by the following factors: (1) soil solid phases (e.g., iron, aluminum, and m a nga nese oxyhydroxides and organic
m a tter) m a y have adsorptive su rfaces that influence soil ch emistry; (2) ad s o rption o r precipitation of m e tals in soils or in roots limits uptake-translocation of most elem ents to shoots; and (3) the phytotoxicity of Zn, Cu, Ni, Mn, As, Bo, Al, and F, for example , limits residues of these elem ents in plant shoots. More recen tly, in ductively co upled plasma-m ass spectrom e try (ICP-MS) and other very sensitive analytical m e thods have been used to examine soil solution and soil-plant transfer of 60 elem ent s as a function of soil pH (Tyler and Olsson, 2001a, b). These studies provide evidence that f u rther supports the conc ept of a soil-plant barrier Several areas for potential errors in the research m e thodology shoul d be avoided when
m aking these comp arisons:

• First, comp arison of application rates is valid only after the system has been equilibrated using accepted m e thods.

• Second, soil pH levels should be equal across rates studie d; co-variance of soil pH should be used to correct for une qual soil pH (Bell et al., 1988).

3-20

### • Third, the m etal concentration in th e source applied affects the slope of metal uptake: higher m e tal concentration in the source m eans higher phytoavai lability at equal m e tal 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 C u, Zn, and Ni. In alkaline soils , the high pH increases uptake of Mo and Se. Pb and Cr are not absorbed by pl ants to any significant ex tent at any pH (Chaney and Ryan, 1993). However, each elem ent should be considered separately because of its uniqu e chemistry. For example, arsenate is more strong ly adsorbed than is arsenite; when a s o il is flooded to grow rice, soil mi crobes can reduce arse nate to arsenite, and the higher concentration of dissolved arsenite can be phyt otoxic to rice in more highly c ontaminated soils. Most other elem ents have little potential for redox change with change in the redox status of soils. Reduced soils can form sulfide, and sulfide form s low-so lubility comp ounds with most of the m e tals of concern in soils, including Pb, Zn, Cd, Cu, and Ni (see above di scussion on sedim e nt chemistry Section 3.1.5). For essential elem ents (e.g., Zn, Cu, Ni), low-solubility species can result in deficiency syndrom es. Upon oxidati on of the soil, sulfid e is quickly oxidized, and the m e tals are return ed to more norm a l equilib rium reactions of aerobic soils.

3.1.7. Atmospheric Behavior/Chemistry Most m etals and m etal com pounds exist in the solid phase under norm a l environm ental conditions and thus occur alm ost exclusively in the particle phase of the atm osphere, where they are ubiquitous. Anthropogenic sources include co m bustion from fossil fuels and m etal industries as well as industrial sources employing specific m e tal comp ounds in specific processes. Som e airborne m e tals (e.g., m a nganese and nickel) m a y de rive largely from crus tal sources (U.S. EPA, 1996a). Richardson (2002) included volcanic erup tions and emissions, entrainm ent of soil and dust, entrain m ent of sea salt sp ray, and natural forest fires as si gnif i can t n atura l m etals emission
sources. For purposes of health risk assessm ent, particle size is important. The aerodynami c size and associated composition of particles determine their behavior in the hum a n respiratory system

. Furtherm ore, particle size is one of th e most importan tparam eters in dete rmining the atm o spheric lif etim e of particle s, which m ay be a key consideration in assessing inhalation exposures as well as exp osures related to e xposure pathways involving deposition o nto soil or water (U.S. EPA, 1996a ). Metals emitted by combustion processes (e.g. , the burn i ng of fossil fuels or wastes) generally occur in sm all particle s or fine fraction, which is often characterized by particle s less th an 2.5 mi crons in diam eter (P M 2. 5 

### ). In co ntras t , the larger sized, coarse m ode particles result from m e chanical disruption, su ch as crushing, grindi ng, and evaporation of sprays, or suspensions of dust from c onstructi on and agricultural operations. Accordingly,
m e tals in coarse m ode particles (i.e ., those larger than approxim ately 1–3 μ m ) are prim arily those
3-21

### of crustal o r igin, such as Silicon (Si) , Al, and Fe (U.S. EPA, 1996a). The fine versus coarse distin ction s i mp ly dif f

erentiates two relative l y distinct size distributio n s o f partic les, the separation point of which occu rs in the range of 1 to 3 μ m . The distinction does not refer directly to particle samp ling m e t hods or size fractionations partic ular to health risk assessm ent (U.S. EPA, 1996a). Fine and coarse particles typically exhibit different behavior in the atm osphere; fine mode particles exhibit longer atm o sphere lifetim es (i.e., days to weeks) than coarse particles and tend to be m o re uniform ly dispersed across a large geographic region (U.S. EPA, 1996a). Relatively lower dry deposition velocities of fine particles contribute to their pers istence and uniformity throughout an air m a ss (U.S. EPA, 1997c). The larger coarse particles (i.e., greater than 10 μ m ) tend to rapidly fall out of the air and have atm o spheric lifetim es on the order of
mi nutes to hours, depending on their size and other factors (U.S. EPA, 1996a). In m o st cases, m e tals do not undergo transfor m a tion while in the particle phase; thus, their rem oval from the atmosphere is regulated by the rate atwhich the particles them selves participate in wet and dry deposition processes. For example, m e tals such as As, Be, Cd, and Pb are generally found in airborne compounds with a single predominate oxi dation state (As(III), Be(II), Cd(II), Pb(II)). S o m e m e tals (e.g., the transition m e tals 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 importantconsideration for health risk assessm e nt as the different oxidation states also di ffer in toxicity (such as for Cr). For m e tals that can change oxidation states , mu c hof the atmospheric chemistry takes place in the aqueous phase, such as cloud droplets or water film s on particles. Metal salts and oxides that dissolve in water can undergo seve ral reversible reactions, including hydration, hydrolysis, polym erization, and r eaction with other anions. T he equilibrium between these for m s depends on the atmospheric conditions, the e quilibrium and solubility constants, and the concentrations of other chemicals. Transform a ti ons between oxidation states can occur either to incre a se th e oxidation state (su c h as oxidiz i ng Cr (III) to Cr(VI)) o r to redu ce it. The s e oxidation or reduction reactions can occur through reaction with other species , such as dissolved m e tals, reduced sulfur species, and organic compounds (Seigneur and Constantinou, 1995). Although models exist that can be used to estim a te m e tal speciation in aerosols with liquid water, the reactions are still highly uncertain. Hg is an exception am ong the commonly occurri ng m e tals; it exists prim arily in the vapor phase under most conditions but can also occur in particle and aque ous phases. The risk assessor should consider at least three species of Hg: el emental Hg, which is larg ely present as a gas; divalent (Hg(II)) inorganic Hg comp ounds, which ar e m ore water soluble; and particulate-phase Hg (Schroeder and Munthe, 1998; U.S. EPA, 1997b)

. The behavior of Hg in the atm o sphere depends strongly on the oxidation state. Elem en tal Hg is capable of being transported long distances, even globally; divale nt Hg deposits within a few hundred kilometers of sources; and 3-22

### particulate m e rcury is deposited at interm ed iate distances, depending on the particle size (Schroeder and Munthe, 1998). Elem ental Hg th at is d e posited can be reemitted from the surface, as can divalent and particulate Hg after biological or chemical redu ction to the elem ental for m . In the gas p h ase, elem ental Hg canbe oxidized to divalent Hg by O 3 

### , OH, H 2 

### O 2 

### , and molecular chlorine (Cl), although other halogen atom s also mi ght be i mportant (Schroeder and Munthe, 1998). In the aqueous phase, elem ental Hg can be oxidized by OH, O 3 

### , and dissolved chlorine, and divalent mercury can be redu ced by processes such as reaction with HO 2 

### and S(IV). Bothvapor-phas e and aqueo u s atm o spheric ch emistry m ay involve mu ltiple phases. EPA has m a de a substantial effort to evaluate the atm ospheric fate of Hg as a result of the requirem e nts of the Clean Air Act. EPA ( 1997b) contains a comprehensive evaluation of
m e rcury’s atmospheric fate, but this is an area of ongoing research and controversy. EPA continues to actively investigate Hg behavior in the atm osphere (e.g., Landis and Stevens, 2003; Jaffe et al., 2003; Bullock and Brehm e , 2002; U.S. EPA, 2001a; Bullock, 2000a, b). Most sampling and analytical techniques publis hed by EPA for m e tals in air are oriented toward evaluation of particular-phase total m e tals rather than m e tal spec ies (U.S. EPA, 1999a). These m e thods involve collection of a particular size fraction of particles (e.g., PM 2. 5 

### , PM 10 

### , and total suspended particulates), wi th subsequent analysis by x-ray fluorescence, atomic absorption, inductively coupled plasm a , proton-induced x- ray emission, or neutron activation analysis gamma spectroscopy techniques. The one notable exception is a m e thod for Hg(Method IO-5) that speciates vapor and particulate form s. To th e exten tthat m etals ar e s o rbed to particulat e phases, analysis of individual m e tal species ca n, atlea s t th eore tica ll y, be accomplished by the
sam e techniques used to analyze th o se species inother so lid m edia.

Potentia l fo r Inhalation Inhalation of m e tals is addresse d in Section 4.2.5.1. Particle si ze is likely to be the most important determinant of potential exposure. Particle size also inf l uences the overa ll f ate of
m etals in the atm o sphere.

3.2. METAL FATE AND TRANSPORT Major m e tal sources to water and la nd include diverse m a nufacturing, mi ning,
com bustion, and pesticide activities. Major at m ospheric sources are oil and coal com bustion,
mi ning and sm elting, steel and iron m a nufacturi ng, waste incineration, phosphate fertilizers,
cem ent production, and wood com b ustion (Haygarth and Jones, 1992). Metals from these atm ospheric sources can find their w a y into soils, sedim e nts, and water. Other m a jor sources to aquatic and terres trial sys t em s in clude chloralkali, acid, pigm ent, electronics, and copper sulfate production.

3-23

### Som e exposure assessments do not involve an thropogenic releases of m e tals to the environm ent. Rather, they focus on changes in exposure to environm ental m e tals that result from other aspects of hum a n activities. For ex ample, acidification of freshwater changes Al phase association and sp eciation, res ulting in an in creas ed do se of naturally occurring m e tals to aquatic biota (e.g., Campbell et al., 1992). Intensiv e irrigation m obilizes Se that is n a turally present in relatively high concentrations inwestern soils, and consequent evaporative concentration in wetlands, impoundm e nts, 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 Pathw ays for Metals When an exposure assessm e nt is completed for a site, the available data are usually limited in their coverage with respect to the ch aracterization of exposure levels over both tim e and space. Depending on the situation, it m a y be a dvantag eous to use a transpo rt an d fate m odel (i.e., a com p utational m odel) to f ill inthe data gap s, such thatan im proved charac teri zation of exposure levels is available for use by the risk as sessor. Models are also useful in situations where it is desirable to have an es timate of future exposure levels th at are expected to result from the implem e n tation of rem e diation measures. These results can be us ed to quantitatively evaluate the effectivenes s of altern ative rem edi ation scenarios that are being considered. The models can also be used to refine the design of the viable alternatives so that the risk assesso r can develop an optim al rem e diation strategy. Although num erous m o dels are available for use, m ost are based on the sam e two funda m e ntal principles: (1) m e tals are ubiquitous in the environm ent, as they are found in the aquatic, terrestrial, and atm o spheric comp art m ents; (2 ) with in each comp artm ent, they are present in as sociation with wa ter (f re ely disso lve dm etal or as organic and inorganic m e tal
complexes), particles (sorbed, precipitated, or inco rporated within a mi neral phase), and air. The evaluation of m etal transport th erefore requires evaluation of th e distribution of the m e tal am ong these phas e s , within each compart m ent (air, wate r, soil), as well as the move m e nt (i.e., the transpo rt) of each of these within andam ong the various com p artm ents. It is im portan tfor the
assesso r’s m odel to s i mu late the m o vem ent of water and p articles exp lic itly because th is provides a w a y to evaluate differences in the de gree to which various chemicals/m e tals m a y be transpo rted in associatio n with partic les ( i .e., via settling and resuspension) or in association with the dissolved phase (diffusive flux of dissolved m e tal). The analyst repres ents th e environm ental setting of interest as a series of discrete, interconnected volumes. Mass balance equations for air, water, solids, and m e tal are then formulated fo r each volum e to obtain a system of mass balance equation s th at m ay then be solved for the con cen trations of interest over both tim e and space. Includ ing all com p art m ents in ev er y model is not necessary. For ex am ple, m odels for a site that is impacted by a sm elter mi ght call for use of a model of an atm o spheric comp art m e n t (e.g., to simu late transport of a releas e from a stack ) and a ter res trial comp artm ent (to s i mu late f ate of 3-24

### atm ospheric inputs to the soil). Alternatively, for an aquatic setting dominated by previously contaminated sedim ents, it m ay be sufficient to consider water and sedim e nt transport alone and to neglect th e atm o spheric and terres trial com p artm ents. Although the underlying principles of m o st m ode ls are similar, the included features vary widely from one m odel to the next. The output from a relatively simp le model m a y be adequate for decision-m aking purposes in som e instances. Som e relativ ely sim p le models are limited in their applicab ility to s t ea dy-stat e ana l yses a nd spatia lly unif o rm condition s, but if this will provide a reasonable and/or conser vative representation of conditions at a site, it m a y provide the analys t with a useful and cost-e ffectiv e m odeling alternative. Inother cases, the analysis m a y require the com p letion of tim e-variable simu lati ons to properly represent conditions that vary over tim e , s u ch as daily or seasonal variations in flow and upstream boundary concentrations, point source loads, and pulse exposures. The an alyst should select an appropriate m odel because not all m ode ls will b e ap plicab le to every situatio n. Although, in prin ciple, the m ore sophisticated m odels provide the risk assessor w ith the capability to comp lete a m o re detailed and m echanistic ally b ase d analys is th an will a simple m odel, successful application of these models will require greater resou r ces (data, time, and funding) than will the use of a simp ler model. The analyst must also possess a relativ ely high level of m odeli ng expertise to use the more sophisticated m odeling approaches. It is for this reason that the mo re sophisticated m odels
are usua lly r eserved f o r use in high e r leve l, def i n itiv e asse ss m ents. Many of the m odels available for use in evalua tin g the transp ort and f ate of m etals were originally developed for application to neutral organic chemicals. As a result, these models frequently include a variety of re actions that are not necess arily re levan tto an analys is of m etal transport and fate (e.g., biodegr adation, photo-oxidation, and vol atilization). Although these models still m a y be of use in an exposure asse ssm e nt for m e tals (the nonapplicable processes often m a y be bypassed), a m ore significant problem is that they often fa il to represent som e important metal-spec if ic processe s. For example, the evaluati on of m etal speci ation and m e tal partition i ng between dis solved and p articulat e ph ases will be repres ented only in a ve ry sim p le
m a nner in such m odels. This limitation m a y be overcom e, at least in part, by performi ng the requis ite m e tal-specific analyses with a sta nd-alone chemical equilibrium model, but this approach will place an added burdenon the analyst to in teg r ate th e resu lts of the two models in a technically defensible m a nner. Although models th at in clude som e m etal-specif i c capab ilities will b e discussed as p art of this Fram ework, no single m odel that is currently available f o r

use inc l ude s all the m etal-spe cif ic features that would be desirable for use. More detailed discussi ons of these and other transport and fate m o dels, as well as a num ber of che mical equilibrium m odels, m a y be found in Paquin et al. (2003) fo r aquatic sys t em s and Allen (2002) for terres trial s y stem s. Addition ally, w o rk is in progress to develop updated m odels that will o ffer improved m e tal-specif i c capabilities (e.g., the

3-25

### Unit W orld model (Di Toro et al., 2001). As a re sult, m etal fate and transport m odels should be viewed as an evolving tool, and new m odels may be expected in the near future.

3.2.1.1. Aquatic Transport Models Modeling of m e tal transport and fate withinaquatic system s invol ves the representation of hydrodynami c transport to simu late m ove m e nt of water, particulate transport to simulate the
move m e nt of particles, and chemical transfers and kinetics to simulate exchange of metal between dissolved and particulate phases and be tween the water column a nd benthic sedim e nt (Figure 3-2). The risk assessor has the opti on of using independent hydrodynami c transport, sedim e nt transport and chemical fate m odels, or an integ r ate dm odel that incorpor ate s all these processes.

Figure 3-2. A generalized model framew ork for chemical fate and transport in an aquatic system. Source: Paquin et al. (2003).
Modeling the m ove m e ntof m e tals thr ough an aquatic system begins with a characterization of the water m ove ment through the system

. The tim e scale for the hydrodynamic analysis should be repr esented in a way that will satisfy the needs of the sedim e nt transpo rt an d chemical fate analyses that are also being perform e d. For example, low-flow 3-26

### conditions associated with mi ni mu m dilution m a y be judged to be the m o st critical conditions in a setting involving a point source discharge, while peak flow conditions m a y need to be simu lated ina setting wh ere resuspension of contaminated sedim ents is the prim ary concern. Asteady-state m odel mi ght be appropriate for use in the fo rm er case, and a tim e-variable model would likely be needed inthe latter case. Thus, the details of th e specif i c p r oblem settin g will nec es sarily have an inf l uence on both m o del se lection and how the m odel will be used. The risk assessor also should be aware of se veral reviews of fate and transport m odels that have been completed since the early 1980s. In earlier reviews, Delos et al. (1984) reviewed models for wastewater load allocations, and Mi lls et al. (1985) desc ribed screening-level analyses for water quality assessm ents of conven tional pollutants. The Agency subsequently prepared an updated summ ary of the features in cluded in the water quality m odels that were available for use in exposure assessm e nt s (U.S. EPA, 1987). Schnoor et al. (1987) simultaneously published their review and includeddetailed descriptions of fate and transport models and their required input param e ters. Late r, EPA published a review describing the use of modeling tools for the developm ent of Total Ma ximu m Daily Loads (TMDLs) in watersh e ds (U.S. EPA, 1997a). More recently, Paquin et al. (2003) completed a review of exposure, bioaccumu lation, and tox i city m odels for aquatic sy stem s, with a focus on their applicability to
m e tals (exclusive of org a nom etallics )

. Because of the advan ces thathav e been m ade since the
mi d-1990s with regard to the de velo pm ent of fate and transp ort m odels, including so m e recent ef

f

o rts to co uple the se models with m etal speciation m odels and m o re sophisticated stand-alone hydrodynamic and sediment transport m odels, the latter review s by EPA (1997a) and P a quin et al. (2003) tend to include the m ostup-to-date inform ation with re gard to the availability of models that are app r opriate for use. These rev i ews also include example applications of m a ny of the m odels discussed. Fate and transport analyses m a y be perform e d by using an integrated hydrodynami c, sedim ent, and chemical transport m o del or by em pl oying what tend to be re latively sop h isticated stand-alone versions of these three subm odels. The advantage of the form er approach is that integration of the hydrodynami c, sedim e nt, and ch emical transport results takes place in a
seamless m a nner with limited need fo r an assessor’s interv ention. This is in contrast to the use of stand-alone m odels, where the output of one model should be for m atted in a way that ensures it is am enable to use withthe othe r models th at are to be applied. A distinct advantage of the latter approach is that it has the potential to reduce the tim e needed to complete a m o del run, an important co nsideration f o r a mu ltiyear simu latio n of a large and comp lex problem setting. For
example, it m a y not be necessary to repeat the simulation of hydrodynamic and sediment transport if the m odel input being m odified affects only chemical tr ansport (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 m odel that include s m etal spec iation ve rs us one that u ses a stand

- alo n e chemical equilibrium model to satisf y th is n eed. The partitio n coef f i cient, which controls th e distribution of m e tal between the dissolved and particulate Anal yt ical S o lution Models 

### phases, is considered to be a key m o del param e ter in Analytical sol ution model s of aquatic system fate and 

### chemical/m etal fate and transport evaluations (s ee Section transpo rt are the simple st 

### 3.1.4). It is important for se veral reasons. First, the model s with the lowest comp utationa l requi reme nts. 

### distribution of m e tal between the dissolved and particulate They are solv ed analytically to provide con c entration s ov er 

### phases has a direct impact on the m a gnitude of particulate metal of time or sp ace. 

### fluxes of m e tal that occur in a ssociation with the settling and resuspension of sorbed m e tal. Second, the part ition coefficient also controls the m a gnitude of diffusive fluxes of m e tal between the sedim e nt in ters titia l water and th e o v erly ing water column, as this flux is proportional to th e concentration gradient of total dissolved m e tal (free ionic m etal + m e tal-DOC and m e tal-inorganic ligand compl e xes) be twe en these co mpartm ents. Partition i n g reactions also affect the m e tal bioavailability and the route of exposure to a m e tal (food vs. water). Given its importance, the risk asse ssor should recognize th at the assum ption of equilibrium partition i ng is a poten tial oversimplification in so m e situation s (e.g., near a point source discharge to a receiving water or imm e di ately 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 m a y be less app r opriate than is othe rwise the cas e and addition al unce rtainty will nec es sarily be ass o ciated withthe m odel results. If this situation exists, consideration should be given to conduc ting special studies (e.g., m easure m ent of free
m etal concentrations ov er tim e in a water sam p le) to test the valid ity of the equilibriu m assum p tion.

Modeling of Hg and m e talloids, such as As and Se, is complicated by transfor m a tion processes that change the for m of the m etal. Similarly, m etals th at readily change oxidation state (e.g., Cr) als o require ad dition a l con sideration s in fate m odel i ng. Many o f the sam e transport models can be used, but input parameters will require m odification. Such m odels are beyond the scope of this fra m ework, which is focused on inor ganic form s of m e tals. See, for example, the Hg cycling model by Hudson et al. (1994) for further guidance on fate and transport models for Hg. All m o deling exercises are limited by the validity of the model Fram ework, the accuracy of input parameters, and the experien ce of the analyst (Dzom ba k and Ali, 1993). Therefore, all exposure assessm ents should include an expl icit description of m odel assum p tions and associated u n certainties.

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### 3.2.2. Terrestria l Fate and Transport Movem ent of m etals thro ugh soils is depende nt o n the chemical properties contro lling speciation, the presence of ligands that control comple xation of m e tals within pore water (and ground water) and adsorption onto mi ne ral surfaces, and the rate of water flux through the soil. Metals are lost from the soil by leach ing into g r o und water and through m etal specific uptak e by plants (see S ection 3.1.6.5). Sections 3.1.4 and 3.1.6 review the processes and m odels that predict m o vem ent of chemicals thro ugh soils o r partition i ng onto mi neral surfaces (i.e., partition coefficients ). These serv e the sam e role as th e fate and transport m odels discussed above for aquatic system s and therefore are no t repeated here.

3.3. ATMOSPHERIC FATE AND TRANSPO R T Fine and coarse particles typically exhibit different behavior in the atm osphere: fine mode particles exhibit longer atm os pheric lifetimes (i.e., days to w eeks) than coars e p articles and tend to be m o re uniform ly dispersed across a large geographic region (U.S. EPA, 1996a). Relatively lower dry deposition velocities of fine particles contribute to their pers istence and uniformity throughout an air m a ss (U.S. EPA, 1997c). The larger coarse particles (i.e., greater than 10 μm ) tend to rapidly fall out of the air and have atm ospheric lifetim es on the order of
mi nutes to hours, depending on their size and other factors (U.S. EPA, 1996a). In the atm os phere, the behavior of Hg depe nds strongly on its oxidation state. Elem ental Hg is capable of being transported long distances, evenglobally; divalent Hg deposits within a few hundred kilom e ters of sources; and particulat e Hg is deposited at interm ediate distances, depending on the particle size (Schroeder and Munthe, 1998). Elem ental Hg that is deposited can be reemi tted from the surface, as can divalent and particulate Hg after biological o r chemical reduction to the elem ental form

. Deposition Because m o st m e tals occur as particul ate m atter (hereinafter referred to as particulates ), their rem oval from the atmosphere is governed by the rates of wet a nd dry deposition. These deposition processes represent an important route of exposure for plants, anim als, and hum a ns. Som e of the deposited metals can com e to resi de on the surfaces of vegetation, resulting in potential ex posure to th e plants and to anim als th at eat the plants. Metals deposited from the atm o sphere also com e to reside in the leaf litt er and upperm ost soil surfaces, where they can be a source of exposure for soil orga nism s, people, and wildlife.

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### 4. HUMAN HEAL T H RISK ASSESSME N T FOR ME TALS

S ourc e : I P CS -E HC #214( 20 00). H u ma n H e a l t h R i s k As se s s m e nt P a ra digm

S ourc e : I P CS -E HC #214( 20 00). H u ma n H e a l t h R i s k As se s s m e nt P a ra digm 

### The National Research Council (N AS/NRC, 1996, 1994b, 1983), of the National Academy of Sciences (NAS), described four phases to the hum an hea lth risk
assessm ent paradigm (Hazard Identificatio n, Dose-resp onse Assessm ent, Exposure Assessm ent, and Risk Characterization) and identified ris k 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(referr ed to as “hazard characterization” in recent EPA docum ents) involves the determination of whether a chemical is or is not causally linked to partic ular health effects. Dose-res ponse involves the determination of the relationship between the m a gnitude of expos ure and the probability of occurrence of the health ef f

ects in question

. A parallel step in the p r ocess toward the hazardidentificatio n and dose-response assessm e nt is exposure assessm ent. In Exposure Assessm ent, the risk assessor quantifies the total expos ure to a toxic agent in the envir onm ent based on am ount taken into the body, including any combination of the oral, inha lation, and derm al routes of exposure. For
som e assessm ents specific to a single exposur e route, exposure m a y be expressed as an environm ental concentration (e.g., ambient air or water concentrations). Depending on the application, the exposure assessm ent m a y be specifi c to a site, a population at a specific location, or it m a y broadly cover a region or an entire nation. Risk Characterization is the final step in the NAS paradigm

. In this phase, th e risk assesso r s u mm arizes and inte rpre ts the inf o rm ation f r om hazard identification, do se-response, and expos ure steps, often by quantitatively comparing exposures w ith doses that are associated with po tential health effects. Risk Characterization addresses the nature and often the magnitude of the hum an he alth risk s, including attendant uncertainty. These steps are addr ess e d in greater detail in the f o llowing section s, withparticular atten tion to the aspec t s s p ecif i c to m etals. The inform a tion provided here comp lem e nts that given by the available A gency guidance for the risk assessm ent process, e.g., for car cinogen risk assessm e nt(U.S. EPA, 2005, 1986), exposure assessm ent (U.S. EPA, 1992c), developm ental toxicity (U.S. EPA, 1991), neurotoxicity (U.S. EPA, 1998c), chemical mi xtures (U.S . EPA, 2000b, 1986), and cumulative risk (U.S. EPA,
4-1

### 2003e), and focuses on the unique and specific ch aracteristics of m eta ls and m e tal compounds that m a y be applied in metals risk assessm ents for hum a n health.

4.1. METALS PRINCI PLES Metals are associated with a variety of health effects that ar e reviewed in detail in EP A’s Integrated R i sk Inform ation System (IRIS) Toxicology Reviews, the Agency for Toxic Substances and Disease Registry (ATSDR) Toxi cological Profiles, the World Health Organization’s International Programme for Chemical Safety (WHO/IPCS) Environm ental Health Criteria Docum e nts, and m e tal toxico logy reviews (e.g., Lukaski, 1999; Hathcock, 1996; Mertz, 1995, 1993; W a llach, 1985). Metals have speci fic attributes that sh ould be considered in all risk assessm e nts. These principles for m e tals risk assessm e nt (see Chapters 1 and 2) apply in various ways to hum an health risk assessm ents, depending on the scale of the assessm e nt (site specific, regional, or national). This section describes applications of the m etals princ i ples to hum a n health assessm ents within the standard ri s k assessm ent fram ework. Specifically, they fall into th e risk assessm ent paradigm as f o llows:
Background levels Exposure Assessm ent Mixtures Exposure and Effects As sessm ent Essentiality Effects Assessm ent Form s of m e tals Exposure Assessm ent (bioavailability) and Toxicokinetics/toxicodynamics Effects Assessm ent (absorption, distribution, m e tabolism , and excretion [ADME] and toxic ity)
Often tim es, hum an health risk assessors start their analysis with a m e tal-s p ecific reference value (RfD/RfC) and/or cancer pote ncy factor that has been deve loped through a process separate from the health risk assessm e nt. The role of the h u m an risk as sessor is to appropriate ly integ r ate th e reference values and potency factor s with the exposure assessm e nt. Thus, the risk assessor need s an understanding of the toxicological endpoints and m echanism s of action that underlie the derivation of these values to ensure that, for example, the appropriate population and lif e stages are addressed, appropriate dietary aspects are taken into consideration, and the appropriate exposure pathways are considered. For m e tals, frequency and duration of exposure, as well as expos ure concentrations and
m etal species, are im portant param eters for the risk assesso r to consider f o r accu rate dose
assessm ents.

4-2

### 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 questi on. 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 exposur e (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 F act ors Influ e nci ng I n te ractio n Bet w ee n Hos t an d C h e m ical

Source : IPCS-EH C #155 (1994)

.



### Assessm ent of hum an exposures to m etals, as with any chemical agent, includes: (1) identifying how people com e into cont ac t with m etals in the environ m ent (2) determining the concentrations of specific f o rm s (speciation) of the metal in spec if ic
m e dial (e.g., soil, water, air, and biota) (3) identifying the pe rtinent exposure m e tric (via consideration of dose-response
assessm ent) (4) estim ating the exposure m e tric (e.g., oral intake, inhalation exposure concentration, blood concentration), w hich m a y involve qua ntifying relationships betw een exposure concentrations and intakes and (5) identifying sources of un certainty and natural variab ility and, where possible, quantifying these in estim ates of exposure.
4.2.1. Background Levels Metals are naturally occurring constituents in the environment. As a re sult 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 A s 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 hum a n h ealth risk assessm e nts, the term “background” ref ers to all existing m etal sourc es e x cept th e tar g et ed source. A particular challenge for the risk
assesso r m a y be assessing the m e tal levels associat ed with th e source(s) o f interest in light of levels derived from natural and other anthropogenic sources.

4.2.2. Bioavailability The term “environm entally availab l e f r action” re f ers to the p o rtion of to tal m etal in s o il, sedim e nt, water, or air that is available fo r physical, chemical, and biological m odifying influences (Lanno, 2003) and represen ts the total pool of m e tal at a given tim e in a system that is potentially able to con t act or en ter an organism

. Of that portion, the bioaccessib l e fraction (BF) is the am ount thatactu ally inte racts at the organ i s m ’s contact surface and is potentially available for absorption or adsorp tion (if bioactive upon cont act) by th e organism

. Bioavailability is th e extent to which bioaccessible m e tals (see Section 2.3) cross b i ol ogical m e mbranes, ex pressed as a fraction of the total am ount of m etal the organi sm is proximately exposed to (at the sorption surface) during a given tim e and unde r defined co nditions. The concept of m e tal bio a vailability includes m e tal species th at are bioaccessible and are absorbed or adsorbed (if bioactive upon contact) with the potential for distribution, m e tabolism , elimination, and bioaccu mulation in the organism

. Metal bioavailab i lity is specific to the m etal salt and particulate size, the recep tor and its specific pathophysiological characte ristics, the route of entry, duration and frequency of exposure, dose, and the exposure m a trix. The m e ta l salt is influenced by properties of the environm ent such as pH, particle size, m oisture, redox potential, organic m a tter, cation exchange capacity, a nd acid volatile sulfid es. Depending upon the
assessm ent, it m a y be appropriate for the risk assessor to consider m e tal bioavailability and
com p are the bioavailable fraction s u sed to es tim at e reference doses (RfDs), or the equivalent, to those m easured in the diet, drinking water, or air. Prediction of toxicity due to exposure to inorganic m e tals is comp licated by wide variations inthe bioacce ssibility and bioavailability of accu mulated m e tals. The for mof the
m etal (chemical species, com pound, m atrix, and partic le size) influences its bio accessibility. In turn, the m e tal form is impacted by properties of the environm ent such as pH, particle size, m oisture, redox Bioav ailabilit y / Hazard 

### potential, organic m a tter, cati on exchange capacity, and acid Relationship 

### volatile su lf ides. Bioavailab ility (how much of t he ingested If two s ubs tanc es were equally toxic at comp arabl e levels of target 

### m etal inter a cts atthe tar g et site) is inf l uenced b y nutritiona l orga n expo su re, the su bsta nce 

### state (deficiency versus excess) , age, sex, physiological state, with the high er intrin si c bioavailabilitywould pose the 

### pathological conditions, and interactions with other g reate s t ri sk. 

### substances present. It is important that the E xposure Analysis de scribes the sam e bioavailable fraction of the
m e tal(s) of concern as that used when estim ating the reference value (e.g., the RfD). For 4-4

### example, m e asuring tota l m etals in d i etary ite m s m ay include insoluble form s of the metal (particularly if soil contamination is presen t on the foodstuff), so effective exposure is overestim ated. There are both direct and i ndirect approaches to address the relative bioavailability of m etals in the environm ent: (1 ) conduct new anim al toxicology studies using the
m e tal for m e ncountered in the site assessm ent; (2 ) use adjunct scientific data to derive an adjustm e nt to the effective dose identified in the anim al study (e.g., data on the distribution of chemical f o rm s of the metal in th e environm ent or at a contaminat ed site); or (3) use a default
assump tion that the m eta l in th e envir o nm ental samp les is th e sam e as that tested. Of the thre e approaches, the first is m ore scientifically s ound. The second option mi ght be available in som e circum stances but is usually precluded by tim e and financial resource limitations, and the third option, is the m ost health-conservative. A f ourth alte rnative conducted f o r site-spec i f i c as sessm ents is f o r the risk assesso r to estim ate bioavailability through so lubility studies or limited bioavailability studies of specific
sam p les from the site. For exam ple, arsenic bi o availability has been es tim ated f o r soils f r om various contaminated sites (Ng et al., 1998; Fr eem a n et al., 1995, 1993) and also through a series of solubility studies of soil from a site contaminated with mi ne tailings (Ng et al., 1998; Salocks et al., 1996 ). Additional exam ples are anim al feeding studies with j uven i le swine for lead bioavailability adjustm e nts or in vitr o tests, altho ugh the Agency currently requires addition a l validation of the latte r ap proaches b ef o re they can be used as the sole basis for m aking bioavailability adjustm e nts (U.S. EPA, 2006a).

4.2.3. Susceptible Pop u lations Risk assessors must specifically consid er population subgroups, which m a y have a greater risk to m e tals thanthe general population (U.S. EP A, 2006b). Factors influencing susceptibility to m e tals include life stage, life style, gender, reproductive st atus, nutritional state, pre-existing health conditions ordisease, and genetic polym orphism s. Chi l dren and elderly people do not regulate m e tal uptake and distributi on efficiently and m a y be at higher risk of
accumu latin g toxic levels (U.S. EPA, 2006b). Pr egnant and lactating wo m e n have a higher dem a nd for essential elem ents, and lack of adequate levels of protein in th e diet can affect the bioavailability of essential elem ents (NAS/NRC, 2000). Individuals with chronic liver or kidney disease m a y have a lower threshold for effects be cause thes e are two of the m ajor target organs of m e tal toxicity. Several well -known, heritable genetic alterati ons affect people’s ability to regulate Cu or Fe, resulting in various deficiency or toxic ity problem s ( W HO/IPCS, 2002). Although many of these sam e factor s are considered in all hu m a n health risk assessm e nts, each has attribu t es specific to m etals-associated risk s.

4-5

### 4.2.3.1. Life Stage In addition to higher intake per kilogram of body weight (Plunkett et al., 1992), children
m a y also be more sensitive than adu l ts to m e tal ir ritants s i nce sensitivity to skin ir ritan t s is generally co nsidered to decrease with age. Infa n t s in th e imm ediate postnatal p eriod can also b e more susceptible to systemic effects of m e tals because absorption of esse ntial m etals is poorly regulated (WHO, 1996). At the o th er extrem e, older adu lts are m o re sensitiv e to m etals th at target the ki dney (e.g., Cd ) because renal function d eclines with age. Efficiency of intestinal uptake of som e trace m e tals, p articularly Zn a nd Cu, also declines as peo ple age (WHO, 1996; IPCS, 1994).

4.2.3.2. Demographics Dif f

erences in lif es tyle inf l uence m e tal expo sure. The risk as sessor shou ld explic itly investigate different lifestyles of the population of concern. Fo r exam ple, the use of dietary supplem ents and other consum er products contai ning essential elem ents has increased. In addition, folk rem e dies such as colloidal silver “cure-alls” and folk rem e dies containing lead tetroxide m a y contain high levels of m e tals (McKinney, 1999; Yanez et al., 1994; Pontifex and Garg, 1985; Trotter, 1985; Bose et al., 1983; CDC, 1983, 1982, 1981; Geffner and Sandler, 1980). Sm oking provides significa nt exposure to som e m e tals (e.g., Cd) and can potentiate the effects of exposures from other sources, and excessive alcohol consumption can exacerbate
m etal ef f

ects.

4.2.3.3. Pregnancy and Lactation Pregnancy and lactation increase dem a nd for so m e essential m e tals, particularly Cu, Zn, and Fe (NAS/IOM, 2003; Picciano, 1996). Because of physiological changes that include higher Fe and Ca requirem e nts, horm onal changes, and su sceptibility to respiratory disease, Z uurbier and Van den Hazel (200 5) suggestedthat p r egna nt wom en could be pred isposed to th e toxic effects of beryllium (Be), Pb, and Mn (2005)

. Recomm ende d dietary allowances (RDAs) specific to p r egnant and lactating wom e n have be en developed for a number of essential elem ents (NAS/NRC, 2000, 2001) an d should be considered by ri sk assessors looking at these
m etals. Addition ally, wo m en lose significant am ounts of Fe during m e nstruation, which m a y lead to increased absorption and toxi city of Cd (Berglund et al., 1994).

4.2.3.4. Concurrent Damage or Disease In general, people with allergies and thos e with pre-existing m e dical conditions have higher-than-average biol ogical sensitivity to environmental stressors. For example, diseases o r treatm ents that dam age the kidney or liver m ay in creas e sens itivity to m etals thattarget these organs. Dam a ge to the lung from s m oking can potentiate effects of simultaneously or
4-6

### subsequently inhaled m e tals, particularly those that act directly on the lung (e.g., Be, Cd, Cr, and Ni). Skin abrasions or other ir ritations also can alter exposures to and subsequent effects of
m e tals (although derm al absorption is not a primary route of m e tals exposure for intact skin).

4.2.3.5. Nutritiona l Sta t

e Risk assessors should be awar e that dietary differences canaffect the absorption of
m e tals, thus changing internal targ et dose. For example, lack of protein (or the type of protein) can decrease m e tal uptak e, thus reducing potential risk. However, relatively little is known about this risk factor and nutritiona l state is an area for further study.

4.2.3.6. Ge netic Poly m o

rphisms a nd Variability Risk assessors should be aware of several well-known, heritable, genetic polym orphism s that affect susceptibility to m e tals. The best known of these are two di s orders that affect Cu
m etabolism : W ilson’s disease and Menkes syndr om e. W ilson’s disease is an autosom al
recess i ve ab norm a lity (p revalence of 1 in 30,000) th at causes impaired biliary ex cretio n of Cu, resulting in accumu latio n in and dam age to various tissu es, particularly the liver, b r ain, kidney, and cornea; hem o lytic anemia also can result

. Menkes syndrom e is an X-linked recessive disorder of Cu m e tabolism (prevalence of 1 in200,000) that resem bles Cu deficiency regardless of level of Cu intake (WHO/IPCS, 2002). He m o chromatosis is another comm on inherited d i sorder. It is characterized by excessive Fe absorption, elevated plasm aFe 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 geneti c polym orphis m affecting Pb m e tabolizing enzym e s is delta-ami nolevulinic acid dehydratas e (ALAD), which has been found to affect the relationship of bone Pb to the cumulative blood i ndex, suggesting that the transfer of Pb from blood to bone is altered. It is suspected that genetic polym orphism s also exist for As m e tabolism (NAS/NRC, 2001), but these have not yet been defined. Risk assessors should consider the possibili ty of genetic differences when assessing potential sensitization reactions. Chronic beryllium disease (CBD) is an immune response, with sensitivity determined by m a jor histocompatib ility (MHC) class II genes (U.S. EPA, 1998d). Similarly, sensitiv ity to Ni is relatedto th e geno type of hum a n leukocyte antigen s (HLA) 3 

### (ATSDR, 2003).

3 Th e m a j o r h isto c o m p atib ility co m p lex is a g r o u p of g e n e s on ch ro m o so m e 6 t h at cod e fo r th e an tig en s that d e termin e tissu e and b l oo d com p at ib ility. In hu m a n s , h i stoco m p a tib ilit y an tig ens are called h u m an leuko cyte antigens because they were originally discovered in largenum b ers on lymp hocytes. The r e arethousands of co m b in atio n s of HLA an tig ens. 

### 4-7

### 4.2.4. Environmental Release, Transport, and Fate The exposure component of a hum an h ealth risk assessm ent includes source characterization and analysis of fate and tr ansport of the chemi cal of interest through environm ental m edia. Models for transport and f ate of m etals from e mission points to environm ental receptors of i mportance to hum ans (e .g., soils, plants, or anim als used in food and fiber) are co vered elsew h ere in this Fram ework docum ent (see Section 3.2), as they are similar for both hum an health and ecological risk assessm ents. It is recomm ended the risk assessor conduct this portion of the assessm ent simu ltane ously for both hum a n health and ecological
assessm ents to ensure ap propriate integration of the results. Hu m an activities thataf f

ect the contact tim e of people with contaminated m edia al so influence the route(s) and total amount of exposure.
4.2.5. Route-Specific D i fferences in Effects Risk assessors should consider how route of exposure affects m e tal bioavailability and whether effects will occur at portal-of-entry or wi ll be due to systemic, target-organ responses. Interactions am ong metals or other exogenous or endogenous compounds also can affect bioaccessib i lity of m etals and are ro ute depe ndent. Thus, many m e tal exposure is sues are specific to the route of entry and will be di scuss e d separately in the f o llow i ng sec tions.

4.2.5.1. Inhalation Exposure Most airborne m e tals, with a few importantexceptions (e.g., Hg and arsine) occur in particulate form . This necessitates certain cons iderations for inhalation exposure assessm e nt, e.g., how particle size af fects respirability (i.e., ho w much of the pollutant enters the respiratory system ). Addition ally, in halation dos im etry f o r particle s invo lves som e distinc tly d i f f

erent processes than for gases (i.e., de position, clearance, di ssolution, etc.), which are also influenced by particle size (U.S. EPA, 2004, 1997c). Particle size is thus an important factor in assessing
m e tals exposure, with the focus generally being on particles less than or equal to 10 mi crons ( F ) in diam eter (PM 10 

### ). Larger particles usually do not pene trate f ar into th e r espirat ory tr act and can be cleared to the ing e stion r oute and swallow e d. Larger partic les m a y have a larger role as an irritant, affecting a person’s eyes and na sal passages, and, if de posited in the uppermost reaches of the respiratory tract, m a y be transfe rredto the d i gestion tract. Thus, for exposure
assessm ents involving m easurem ents (e.g., using area orpersonal sam p les), the particle size is an important factor in determining inhalation exposure to m e tals. Since inhalation is a prim ary route of exposure for m e tals, the risk assessor should have a good understanding of inhalation dos im etry m ethods and how input s vary for m e tals. Key
m e thods for inhalation dosim e try are described in EPA guidance docum ents (U.S. EPA, 2004, 1997c) and a num ber of models are available for cal culating relative region al respiratory tract
4-8

### deposition in rodents and hum a ns (reviewed in U.S. EPA, 2004). The guidelines for reference concentration (RfC) developm ent (U.S. EPA, 1990) cite the regional deposited dose ratio (RDDR) m o del, which has been used for developm ent of a number of RfCs for m e tals. The multipathway particle do sim etry m o del (MPPD) d eveloped by the Chemical Industry In stitu te f

o r Toxicolo gy (CIIT) an d the Dutch National Ins titu te of Public Hea lth an d the Enviro nm ent (RIVM) was used in EPA’s Air Quality Crite ria for Particulate Ma tter (U.S. EPA, 2004). MPPD i mproves lung dose estim a tions by considering life-stage-specific param e ters, particle clearance from the lung, and differences in oron asal breathing patterns with work load. The hum a n equivalent concentration (HE C ) is the con centr ation th at is b elieve dto resu lt inthe sam e dose to the respiratory tract re gion of interest as was receiv ed by the exp erim ental an im al species. In developing inhalation exposure estim a tes, the risk assessor should pay careful attention to the form of the m etal pertinent to the dos e-response assessm ent (e.g., RfC, IUR). Simply
m easuring the total am ount of a m et al without regard to speciation m a y introduce uncertainties into inhalation exposure estim ates or other exposur e routes. Metal speciation affects a range of processes that change how the m e tal is deposit ed 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 assessm ent includes that specif i cation

. The bioav ailab i lity of m etals via in halation can be much higher than that of other intake routes. This m a y result in relatively high internal dos es, even when intak es are similar to those from other routes. An exa m ple is the large contribution m a de by cigarette sm oking 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) m a y alte r the deposition pattern of inhaled particles and contribute to variations in bi oavailability (Jam es, 1994; Xu and Yu, 1986; Phalen et al., 1985). For m o re inform ation on the consideration of particle size in the dose-resp onse asses s m ent for RfCs and IURs, the risk assessor should refer to U.S. EPA, 1990. For m e tals having alternative Agency-developed dose-response m e trics (e.g., blood Pb concentration), respirability, deposition, and clearance as well as absorption in to the circulatory system m a y need to be addressed as part of th e Exposure Assessm ent. Risk assessors should be aware of emerging issues in inhalation dosim etry that m a y have important impacts on exposure assessm ents for m e tals (U.S. EPA, 2005). The developing literature suggests that current dosim e try m ode ls and traditional dose m easures (such as concentration in m a ss/unit volum e) m a y not adequately characterize hu m an health risk to ve ry sm all partic les, such as p article s < 2.5 μm (PM 2. 5 

### ) or <1 μm in diam eter (nanoparticles). Much of the recent w ork on nanoparticle deposition has been conducted with m e tal oxide particles (e.g., titanium dioxide), and a growing body of literature is becomi ng available to the risk as sessor. The bioavailability of Pb and other m e tals appe ar s to inc r eas e with decr ea sing partic le size, 4-9

### particularly from the inhalation a nd oral routes of exposure, so ri sk assessors should be aware of the poten tial im plication s for bioava ilability of metal n a noparticles. Risk assessors should also consider exposur e to m e tals in shower water, in which aerosolization can occur from the hot water tap. Though the m a gnitude of exposure from showering is unknown and comparable m odels do not exist for aerosolized m e tals (W ilkes, 1998), m ode ls have been developed to predic t hum a n inhalation exposures due to volatile organics from showering (e.g., Guo, 2002; Moya et al., 1999; McKone, 1987). Wher e appropriate, the risk assessor should address such inhalation exposur es during the Risk Characterization phase.
4.2.5.2. Dietary Exposure Risk assessors should be aware that dietary pathways represent a m a jor exposure route for Dieta r y Exposure



### m e tals (Choudhury et al., 2001). Estim a tion of Due to the di versity of the human di et, there may be wide vari abilit y in the intake 

### intakes of metals occurring in food requires of some meta ls in food (e.g ., seafood ), resulting in both temporal variability (e. g ., 

### inform ation on the levels of m e tals in food and the spi k e s after a seafoo d mea l ) and 

### am ount of food consumed (NAS/IOM, 2003). A g eo g ra p hic or cultural variabilit y
. 

### num ber of references provide assessors with national-scale inform ation on dietary exposure to m e tals (Capar and Cunningham , 2000; Schoof et al., 1999a, b; Thom as et al., 1999; Bolger et al., 1996; Dabekaand McKenzie, 1995; Gunderson 1995; Tsuda et al., 1995; Dabeka et al ., 1993). Although large-scale surveys of the
m e tal contents of foods and food consumption patterns have been conducted (e.g., Egan et al., 2002; Ryan et al., 2001; U.S. FD A, 2001; O’Rourke et al., 19 99; Thom as et al., 1999; U.S. DHHS, 1996), assessors should be aware that these surveys have se veral limitations f o r applications to hum an he alth risk assessm ent. Analysis is often conducted with “m arket basket”
samples of packaged processed f oods. W ith a few exceptions, such applications have not been
empirically evaluated against biom arkers of exposure (Clayton et al., 2002, 1999; Choudhury et al., 2001). Risk assessors should be mi ndful that food consumpti on surveys are generally limited to short-term consumption (e.g., 1-3 days) and do not capture intra-indivi dual variability that would affect long-term averages. Furtherm ore, dietary patterns m a y change over tim e (e.g.,
consumption of ethnic foods in childhood m a y change later), and, thus, patterns discerned at any given tim e m a y not accurately rep r esent his t oricalor future exposures. An addition a l challenge facing the risk assessor is integration of data from separate m e tal residue and food consumption surveys (e.g., Tom e rlin et al., 1997). This leads to considerable un certainty in estim a tes of m e tal exposure via the dietary route.

4-10

### 4.2.5.3. Incidental Soil Exposure Infants and children can have enhanced exposures to m e tals through the pathway of surface dust because (1) they crawl an d play in cl o se proximity to surface du st and (2 ) they often mouth their hands (e.g., finger sucki ng) and objects in their environm ent. This causes an intake of surface dust that is g e nerally greater than that which is norm a lly 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 m a y a c t to reduce overall dus t intake. However, ri sk assesso rs sh ould be aware thatd ata ar e limited with reg ard to dis tinguishing between the quantity of dust ingested and the quan tity of soil ingested. Th is param e te r is important in connecting m easuredsoil m e tal concentratio ns with surf ace dust ing estion that o ccurs in the indoor and o u tdoor env i ronm ents (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/sup erfund/program s/risk/rag sa/ind e x.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 partic ipate in activities in which inte nsive exposure to surface dust and soil occur (e.g., occupational gardening, construc tion work) are lower than those of children (Calabrese et al., 1990; Hawley, 1985). Because con centrations of the metal contaminants in soil can be expected to vary with depth, risk assesso rs should consider soil me tal concentrations at the depth appropriate to the me tal(s) of concern as well as human behaviors and activities.

4.2.5.4. Drinking Water Exposure Treatm e nt of surface and/or ground water fo r hu m an consumption rem o ves disso lved organic carbon and suspended orga n i c sed i m ents that c an f o rm com p lexes with m etals (AWW A, 1999). Thus, inorganic form s of m e tals in treate d drinking water will ofte n consist of the m o re bioavailable, water-soluble specie s. Treatm e nt also rem oves bact e r ia that can partic ip ate in organification reactions of toxi cological significance to hum ans (e.g., m e thylation of inorganic
m ercuric m e rcury). Risk assesso rs estim ating the in take of m etals in drinking water will require inform ation about concentrations of m e tals in the water and the am ount of water consum ed. Data on the
m e tal content of tap water can be obtained from EPA’s Office of Dr inking W a ter. EPA’s Exposure Factors Handbook contains exposure inform ation on daily drinking water ingestion and incidental ingestion of water during swimmi ng and showering (U.S. EPA, 1997e ). Generally, water m e tal concentrations are m easured at the d i stribution p o int for municipal w a ter delivery system s. Distribution sy stem s within hom e s (pipes, storage containers, etc.) can contribute significant am ounts of m e tals (e.g., Pb, Cu) to the hom e drinking water

4-11

### (Graziano et al., 1996); consequently , the contribution of m e tals from hom e -based pipes, etc., is a source of uncertainty in the hum an health risk assessm ent.

4.2.5.5. Dermal Exposure Metals absorption throu gh the skin is limited because the derm al route of exposure is of less concern during a health risk assessm ent. Ho wever, some m e tals (e.g., Ni and Cr) have the potential to induce toxic and se nsitization effects directly on the skin (U.S . EPA, 2001, 1992). Derm al exposure can also lead to intakes via other routes, such as oral exposure via hand-tomouth transf er or ocular contact. Potential sources of derm al uptake that the risk assesso r sho u ld consid er include sm a ll particles in contact with the skin; m e tal exposure during bathing, showering, and swimming (NAS/NRC, 2002); and the uptake of m e tals through dam a ge d skin (e.g., irritated skin, sunburn). Derm al contact with m etals in so il also repre sen ts a potential route of ex posure, but the relatively low lipid solubility of most m e tals limits absorption through the skin (Paustenbach, 2000; Hostynek et al., 1998). F e w studies have actually attempted to quan tify the extent or kinetics of the derm al penetration of m etals deposited on the skin, and th e applic ability of these studies to
m etal specie s and com p lexes thatoc cur in su rface dust or soil is highly uncertain.

4.2.6. Integrated Exposure Approaches to integrating exposure across pathways and physiological routes of uptake include m odeling, estim ates of relative bioa vailability, and the use of biomarkers.

4.2.6.1. Modeling Risk assessors have access to only a few specifi c, integrated exposure models for metals. The Integrated Exposure Uptake Biokinetic (I EUBK) model for Pb in children (U.S. EPA, 1994a; White et al., 1998) was specifically develope d for translating exposure measurements into risk estimates. The IEUBK model and bac kground documentation are available on line at : 

### http://www.epa.gov/superfund/lead/products.htm

. 

### Risk assessors should not apply this model to other m e tals because it was derived using Pb-specif i c inf o rm ation and consequ ently is pe rt inent only to Pb. A multipathway exp osure model specific for As has also been develope d (Cohen et al., 1998). Less complex models linking adult exposures and blood Pb concentrations are available (C arlisle, 2000; Stern, 1996, 1994; U.S. EPA, 1996b; Bowers et al., 1994; Car lisle and W a de, 1992), and a stochastic hum an exposure model for Pb that is linked to a lead ph arm acokinetics m odel may also be of use to the risk assessor (Beck et al., 2001; O’Flaherty, 1995). Other m odels available to risk assessors are EPA ’ s Total Risk Assessm e nt Methodology (TRIM), which is b eing developed f o r mu ltipath w ay risk ass essm ent f o r air po llutants includ ing
4-12

### metals ( http://www.epa.gov/ttn/f era/trim _gen.ht ml ); EPA’s Stochastic Human Exposure and Dose Simulation (SHEDS) model, a probabilistic, physiologically based model that simulates aggregate human exposures and doses for variou s population cohorts of interest (Dang et al., 2003; Zartarian et al., 2000); EP A’s Dietary Exposure Potentia l Model (DEPM) (Tomerline et al., 1997); RESRAD, a generic exposure model deve loped 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 m a y find biom arkers of exposure, ef fect, and suscep tibility useful as basic tools (IPCS, 2001, 1994). Centers for Disease Control (CDC) and the Nati onal Health and Nutrition Examination Survey (NHANES) are currently developing a national database to quantify and characterize body burdens (based on hum a n blood and urine su rveys) that in cludes Pb, Hg, Co, uranium (U), antim ony (Sb), barium (Ba), Be, Cs Mo platinum (Pt), tha llium (Tl), and tun g sten (W ) (CDC, 2005). Risk assessors can use this survey as a baseline m e asure against which the levels in receptor population individu als can be comp ared. The data are summarized in age, gender, and ethnicity categories. NAS also has completed a substantial amount of work inthis area (NAS/NRC, 2006). Integration of exposures across m e dia, route, and tim e of exposure can be reflected in biom arkers of exposure. A biom arker of expos ure is a m easure of cumu l a tive exposure to a
m etal and also of m etal actually existing in the body, as occurs with chronic exposure to m e tals. However, such an approach m a y not be appr opriate f o r m etals that ar e not extens ively bioaccumulated in tissu es, and it do es not differen tiate betwe en m etal present in a tiss ue in a sequestered or inactive f orm and m e t a l engaged in toxic or pathological processes. The approach also does not differentiate naturally occurring ex posures from those due to added m e tals. 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 histo ry, could lead to an overestim ation o f exposure if the risk assessor does n o t accoun t for seafood consumption (NAS/NRC, 1999). Thus, use of biom arkers increases the need for
com p rehensive, mu ltipathway assess m ents of exposure. Wh en available, ref erence o r baseline levels of biom arkers of exposure shouldbe incorporated in to the assessment. Sour ce: H e n der so n etal . ( 1 9 89) ; I PC S- EH C #15 5 ( 1 9 94)

. H y p o t he tical R e latio n shi p s A m o n g B i o m ar kers of Ex pos ur e



### 4-13

### 4.3. HAZ A RD CHARACTERI Z ATION Hazard iden tification (o r hazard ch aracterization ) is “th e determination of whether exposure to an agent can cause an increased incide nce of an adverse health effect, such as cancer or birth defects, and characteriz ation of the nature and strengthof the evid ence of caus a tion” (NAS/NRC, 1994b). This includes identification of the target organ(s), consideration of any route-specific issues, evaluation of the advers ity o f the effect s observed, and consideration of relevance to hum a ns. Key point s andm e tals-specific concep ts to be cons id ered in h azard characterization are d e tailed in the following sections.

4.3.1. Mixtures and Interactions In m o st settings, individual m e tals exist as comp onents of mi xtures with other m e tals and/or organic substances (ATS DR, 2004; NRC, 1988). Effe cts of the metals in mi xtures m a y be synergistic, additive, subaddi tive, potentiating, and/or an tagonistic. Interactions am ong m etals occur by com p etition f o r binding locations on specif i c e n zym es or on cellu lar re ceptors during the processes of absorption, excretion, or sequestration at th e target site. The presence,
am ounts, and interactions of all inorganic m e tals are important when c onsidering and evaluating the effects of exposure and resulti ng hum a n health risk assessm e nt. Metals usually exist as com ponents of mi xtures with other metals and/or organic substances (ATSDR, 2004; NRC, 1988). Becaus e the inform ation o r guid ance on risk s of m etal
mi xtures is limited, risk assessors should follo w published guidance for the hum an he alth risk
assessm ent of che mical mi xtures in genera l (U.S. EPA, 2000b, 1986) and cumulative risk guidance (U.S. EPA, 2003e). Only a few controlled studies ex ist on the interactions of m e tals relevant to levels found in the environm ent (see ATSDR, 2004, on mi xtures of (1)As, Cd, Cr, and Pb; and (2)C u, Pb, Mn, and Zn). Risk assesso rs m a y use the curren tdef ault app r oach , which assu m e s additivity of the doses for each m e tal as it will p r oduce estim a tes that are overly-cautious. This approach, which involves calculation of a hazard index (HI), is most approp riate for chemicals that p r oduce th e sam e effects by similar modes of action Metal-Bindin g Proteins Metallothioneins (Ag, Hg, Cu, Bi, Cd, Pb, Zn) t 

### (MOAs). However, differing potencies of Tran sfe rrin (F e, Al, Mn) t 

### m e tals with similar MOAs should be accounted Ferritin (Fe, Cd, Zn, Al, B e ) s Cerulopl asmi n (Cu, Fe ) tr 

### for by converting chemical concentrations into Lead -bin ding protein ( s) (Pb ) s 

### an equitoxic dose using either toxic units (TUs) Membrane ca rrie r protein s t 

### or toxicity equivalence factors (TEFs). In the s = sto r age; t = trans p ort; tr = tran sform 

### case of chemicals with different M OAs, the ris k
assesso r sho u ld consid er estim ating s eparate eff ect-spe cific HIs for eachchemical inthe mi xture
4-14

### using the RfD as the tox i city value for each e ffect. Newer EPA guidance provides a n u m ber of quantitative approaches for charact erizing mi xture risks (describedin detail in U.S. EPA, 2000b, 1986). The term s molecula r mimicry and ionic mimicry have been applied to situations in which
a m e tal for m s a comp lex with an endogenous ligand and the resulti ng compound mi mi cs the behavior of a norm a l substrate, disrupting nor m a l function (Ballatori, 2002; Clarkson, 1993). Molecular or ionic mi mi cry m a y be viewed as a fo r m of metal-m etal inter action. Most examples involve the replacem ent of an essential m etal with a nonessen tial m etal. For exam ple, Cd can
mi mi c and substitu te for Zn and Ca. Additiona lly , m a ny diffe rent pro t eins in the body complex with m e tals which m a y modify their toxicity an d kinetics (e.g., som e m e t a ls bind with albumin for purposes of transport in the circulatory syst em and across cell m e m b ranes or within cells). Som e proteins have different bindi ng kinetics for the various m e tals, r esu lting in sp ecific proteinm etal inter a ctions. Risk asse ssors should be fami liar with th ese m e tal-binding proteins to
corre ctly interpre t the b i oavailability of individual m etals within a mi xture and the po tential us e of protein expression as a bi om arker of m etal exposures. Many of the inte ractions between ess entia l m etals are r elated to m aintain in g optim al nutritional levels by synergism s and antagonism s at both physiological an d extrinsic (dietary) sites. W orld Health Organization (WHO ) publications (IPC S , 2002; WHO, 1996) have summarized these hom eostatic interactions, which are often comp lex (e.g., excess C ain the diet
m a y induce signs of Zn deficiency, even if the Z n intake is n o rm al). Similarly, exc es s Zn in the diet m a y aggravate Fe d e ficiency. Interactions between essentia l and nonessential m e tals are very comm on (e.g., Cd uptake
can mi mi c that of Zn). Similarly, among anions , mi mi cry of the sulfate and phosphate ions
occurs. However, the risk asse ssor should be aware that validated data in hum ans are rare, and that applications of this phe nom enon are best limited to screening-level assessm ents. Two nonessential m e tals m a y comp ete for passive tr ansport across common sites on a m e mbrane or with an esse ntia l m etal on an active b i nding site. These effects m a y not be additive and likely relate to relative b i nding streng th. One m e tal m a y affect one site and another m e tal may affect a different site; this can include both active and pa ssive transport or binding sites and may or m a y not inc l ude intera ction s with essen t ial m etals. These ef f

ects of ten will be additive. T h e risk
assesso r sho u ld be f amiliar with the MOAs of each m etal of concern to d evelop atlea st a gene ral understand ing of whether the mi xtu r e effect is li kely to be additiv e or more than add itiv e. Metals can b e active at m o st cellular s ites wh ere organic toxicants ha ve their effects and directly interfere with receptor activation (Stoi ca et al., 2000), ion cha nnel regulation (Kiss and Osipenko, 1994), cell signaling (DeMoor and Kor opatnick, 2000), cell adhe sion (Prozialeck et al., 2002), and gene transcrip tion (Meplan et al., 2000). T hus, m e tals are not readily distinguished from organic substances in the rang e of their p o tential m echanism s of action atthe cellu lar and molecular le vel, so th e risk assesso r can transfer knowledge about toxicity across 4-15

### chemicals. Additionally , co-occu rrence of m e tals with organic substances can change the bioavailability and increase or decrease absorp tion. For example, in the diet, citrates and histidine are known to enhance Zn absorp tion, w hereas ascorbate can m odify Fe-Cu antagonism s . Low protein content may increas e the absorption of Cd and Pb, and oral contraceptives m ay influence the m etabolism of Fe , Cu, and Zn. The risk assessor should be aware of these conditions to avoid over- and underestim a ting risk.
4.3.2. Essentiality Essentiality 

### Certain elem ents are nutritionally essential to Esse ntiality should b e viewed as 

### hum a ns and play a key role in physiological or part of the overall do se -respon se 

### biochemi cal processes (NAS/IOM, 2003; IPCS, 2002; relation shi p. The sh ape of this relation shi p can vary amon g 

### WHO, 1996). Elem ents esse ntial to other organism s orga nisms. F o r a given su bpop ulation, referen c e do ses de sig ned t o prote c t 

### m a y not be essential to hum a ns and vice versa. from toxic i ty of exc e ss should not be 

### Adverse nutritional effects can o ccu r if essentia l m etals set belo w do ses ide n tified as e s sential for that sub p o pulation. 

### are not available in suffici ent am ounts, and nutritional deficits also can be a dverse and increase the vulnerability of hum a ns to other stressors, including those associated with other m e tals. Essentiality should be viewed as part of the overall dose-response relationship, and referen ce doses designed to protect from toxicity of exce ss should not be set below doses identified as essential. Meta ls thatare cu rren tly deem ed nutritionally es sential for h u m ans are Co, Cr III, Cu , Fe, Mg, Mn, Mo, Se and Zn (Table 4-1). Som e m e tals (e.g., B, Ni, Si, V, and perhaps As), while not essential to hum a n health, m a y have som e be nefi cial effects at lo w levels of exposure (NAS/IOM, 2003). Risk assessors should consider the essential elem ents as comprising three groups: those that are cations (Z n, Fe, Cu, Mn, Cr), those that are anions (Mo, Se), and those that are a bio-inorganic comp lex (i.e., the Co complex, cobalami n). The hom e ostatic
m echanis m s differ for each group, an d risk assessor s can use this knowled ge to generally classify the types of health effects that mi ght occur and the potential bioa vailability of the m e t a l. In general, the gastrointestinal tr act and the liver regulate the upt ake and transfer of cations (e.g., Fe, Zn, Cu). The anionic group is m o re water-solubl e and is less reactive with N, S, P, O, and hydroxide groups than are cations. They are abso rbed very efficiently through the intestine and their subseq uent com p artm entalization and excr etion is by m a nipulation of their oxidation and
m ethylation states; total body burde n is regulated by renal excreti on. The risk assessor should be aware that hom e ostatic controls do not typically apply to effects at the portal of entry, so effects can be seen at lower doses than t hose 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 (inadequ ate) doses to hig h (toxic) doses (s ee tex tbox) when determining an acceptable upper 4-16

### Table 4-1. Metal essen t iality for humans Nutritionally essentia l 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 Antim ony III, V Arsenic III, V Barium II Beryllium II Bismuth III, V Boron III Cadmium II Cesium * I C h romi um V I Gallium * III Germ anium* 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, I II Tin II, IV Titanium IV Tungsten VI Uranium IV, VI Vanadium III, V Zirconium * IV

* Limited hum an data for these m etals.

exposure limit. Several agencies have developed guidance for selecting a benchm ark dose that is not too low (and, theref ore, likely to re sult in deficiency) or too high (and likely to result in toxicity to
som e segm e n t of the population). IP CS (2002) guidance describes the use of an “Acceptab l e Range of Oral In take” (AROI), which estim ates the mi nim al requirem ent to preven t d eficiency an d 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 R DAs (now known as RDIs) using the estim a ted average requirem ent (EAR) or adequate intakes (AIs). They also de veloped a tolerable upper intake leve l (UL), which is the h i ghe st lev el of daily nu trient intake that is likely to pose no risk of adverse health effects to al m ost all individuals in the gene ral population. The UL is based on a risk assessm ent m odel similar to that used by E P A to set the RfDs and is intended to protect the population from adverse health effects resulti ng from excess exposure to a compound. ULs are available f o r

all the ess e ntia l m etals and f o r B, Ni , and V. ULs do not take into account sensitive Obser v e d Le v e l of In take



### 4-17

### or immuno- compromi se d population s . ULs m a y diffe r from RfDs because they are derived from hum an studies rath er than anim al studies and us e sm aller unc ertainty f actors. Additio nally, the risk assessor should be cauti oned that RfDs are intended to cover sensitiv e 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 com b inations (N AS/IOM, 2003).

4.3.3. Forms of Metals Unlike organic chemicals, m e tals are neith er created nor destroyed by biological or chemical processes. Ho wever, thes e processes can transform m e tals from one valence state to another and can convert them be tween inorganic and organic for m s. Inform ation developed for one for m of a m e tal m a y not be directly applicable to other fo rm s. Di ffer e nt valence states (species) of the sam e m e tal aff ect bio access i bility and bioava ilabi lity, and they elicit different responses in the hum an body. The particle size and environm ental m atrix (water, soil, air), within which the m e tal is em bedded, influence e xposure amount, rate, and r oute, particularly for the inhalation pathway, which can then result indifferent target organs and response levels. Therefore, the risk assessor should consider which fo r m of t he m e tal of interest is being assessed. If exposure or effects inform ation has been deve loped for a different m e ta l species, the risk
assesso r sho u ld either m ake approp riate ad just m e nts or acknowledge th is as a sign ificant uncertainty in the Risk C haracterization. For example, Cr(III) is es sential in the d i et, whereas inhaled Cr(VI) is carcinogenic. Risk assessors should be aware that infor m a tion developed on the health effects of one for m of a metal m a y not be direc tly applicable to other forms, particularly organom etallics; other guidance do cum ents should be consu lted when co nducting ris k assessm ents for organo m etallics.
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 perspec tive, net accumulation in human tissues may or may not be relevant in the hazard ch aracterization of metals (see, for example, Section 4.3.4.2 for a Toxicokine t i cs Tox i co kinet i c s de sc ribe s t he se rie s of processe s t hat dictate the disp ositio n of a sub s tan c e i n or on the body after exposure o c cu rs and pro c e s ses rel ated to depo sition on the body su rface s (al s o co nsid ere d in exposure a ssessment s), a b so rption, distrib u tion, metaboli s m, and e x c r et i o n ( ADME )

.



### 4-18

### discuss i on o f m etal sequ estration an d subsequen trem obilization). Furth e r, ther e ar e n o available simple m e trics that allow quantification of th e potential for hu m a n bioaccumulation of m e tals, although a f u ll pharm oc okinetic m o del can be used to estim ate m etals bio accu mulation an d distributio n in hum an Bioacc umulation of meta ls is the net 

### tissue s. Addition al disc ussion of the potential f o r accumul a tion of a metal in the tissu e of interest o rthe wh ole org anism that 

### rem obilization of sequestered m e tals is in section 4.3.4.2. res ults fro m a ll environm en tal 

### All these processes can be described through the use of exp osure m e dia , inclu d ing air, wate r, solid p h a s e s (i.e., soil, sedi ment), and 

### physiologically based pharm acokinetic (PBPK) models. diet, and that rep r e s ent s a net mass balan ce b e tween upta k e a nd 

### Integrated descriptions provide m e trics of internal dose elimination ofthe metal (SAB, 2006). 

### (including biological m arkers of exposure) that can be used by risk assessors to impr ove the quantitative basis of doseresponse assessm ents. Unique features of m e tals that result in differe nces in toxicokinetic behavior of m e tals as compared to or ganic substances are shown in Table 4-2.

Table 4-2. Summary o f major d i fferences in kinetic beha vior of orga nic compounds compared to metals and in organic metal compounds in humans Organics Metals Tissue uptake is m o st co mm only a blo o dflowlimit ed process, with linear portioning into tissues. Metals and their co mp lexes are often ioni zed, with tissue uptake (m em brane transport) having greater potential to be diffu sion-limited or to use specializ ed transport processes. Metabolism is generally extensive and often speciesspecific. Metabolism is usually l imi ted to oxidati on state transitions and alky lation/dealky lation reactions. Persist ence i n body fat is co mm on because of lipid solubility (not capacity -limited). Often sequest ered , bound t o specific plasm aor tissue proteins (intrinsically capacity -limited) or bone. Due to comp lex metabolism , organics may be eliminated by excretion in urine after biotransformation from lipophilic form s to hy drophilic f o rms, in bile after conjugation to large organic m o lecules, or in exhaled air if n o t metabolized. Predominantly eliminated in urine because metal co mpounds are generally s m all m o lecules and are hy drophilic. As a result of protein bi ndi ng, m a y be excreted via hair and fingernails.

Generally substance-spe c ifi c homeostatic mechanisms are not available. Essential met a ls have homeostatic mech anisms that maintain optimu m tissue levels over a range of exposures. Interactions with other structurally simi lar co mpounds may occur, especially d u rin g metabolism

. Interactions am ong m e tals and between metals and organics are nu m e rous and occur co mmonly during the processe s of absorption, excretion, and sequestration.


### Source: Adapted from Golub et al. (2004).

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### 4.3.4.1. Ab sorption Absorption is a process by which an admi ni stered substance enters the body. Bioavailab i lity is a term often used to describe the degree of absorption. Two elem ents of the absorption process are critical for evaluating systemic doses of m e tal comp ounds, both the degree and rate of absorption. Although inform a tion on both of these param e ters is ideal for developing a quantitative estim a tion of systemic doses, inform ation on the degree of absorption is m o re commonly available for m o st chemicals. Metals have a num ber of unique properties that im pact their absorption across bio l og ical m e m b ra nes. Key factors affecting the absorption of
m e tals include solubility, particle size, valenc e state, lipophilicity, and the exposure matrix. Soluble form s of the metal are m o re read ily abs o rbed sinc e the m etal ion itse l f is typ i cally the absorbed en tity. The bio a vailabil ity of m etals in creas es as p artic le size d ecreases. However, in the lung particle size also determines the site of deposition and thus the clearance m e c h anism s that can ultim ately result in systemic uptake (v ia transpo rt to the lymph system f

o llowing
m acrophage engulfm ent ) or the GI tract (via mu cociliary transport). Chemical speciation in term s of valence state can affect absorption. Re cent progress in identifyi ng m e tal transporters suggests that generalizations are no t appropriate, and each m etal s hou ld be assessed in term s of its ab ility to access tran s porters and the presen ce of transporters in potential target o r g a ns. In general, lipo philic comp ounds will b e absorbedmore readily than hydrop hilic ones. For
example, hum an skin is not very perm eable, and it provides a good barrier against derm al absorption of m e tals and m e tal comp ounds; el emental Hg and dim e thyl Hg are notable exceptions. Risk assessors should note the compl e xities in absorption pro cesses since they have direct implications on m e tals risk assessm ent, prim arily in requiring deta iled consideration in extrapolating across different exposure conditions or anim al species. Absorption can vary dram atically for different for m s of the sam e m e tal, f o r the sam e f o rm of m e tal in dif ferent
m a trices, among different species, and across differ ent routes of exposure. Therefore, it is not appropriate to assum e concordance in absorp tion of m e tal compounds without a detailed evaluation (and docum e ntation) of the scientific basis for such an assump tion. For example,
empirical infor m ation on derm al absorption of metals should be consulted when available (Stauber et al., 1994; W e ster et al., 1992; Hursh et al., 1989; Il yin 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 potentia l target organs and the subsequent target tissue doses. The distri bution of m e tals reflects their transport and
accumu latio n in the bod y within tissues, blood o r plasm a , or other ex tracellu lar sp ace. Partitioning to blood and cellular co mponents, particularly via intera ctions with proteins, is of particular importance for m e tal risk assessm ent. Retention in tissues of m e tals or metal 4-20

### compounds generally is related to formation of inorganic comple xes 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 ina nontoxic form allows the metal to reside in the body without producing a toxic or pathological effect. For example, As a nd 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 re tained in soft tissues (e.g., liver and kidney) for 10 to 20 years by intracellular binding with me tallothionein. Metal binding to proteins is capacity-limited, and toxicity to target organs occurs when the binding capacity is exceeded. Thus, the potential for toxicity ex ists in older adults for metals with long half lives that are initially adequately sequestered. Conversely, rete ntion in tissues is a dynamic equilibrium and can be a source of internal expos ure long after the external expos ure source has been removed. For example, Pb in bone can be mobilized duri ng pregnancy and lactati on or as a result of osteoporosis (USEPA IRIS, http ://w ww.epa.gov/iris/subst/0277.htm )

. The risk assessor should consider all the aspects of me tal 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 informati on to complete a quantitative analysis of these processes during the Risk Characterization.

4.3.4.3. Metabolism Metabo lism of m etals is limited to ox idation-reduction reactio ns or alkylation/ dealkylation reactions. In these reactions, new inorganic species or m e tal organic complexes
m a y be for m ed, but the m e tal ion persists. Ne vertheless, d i fferences in th ese tr ansf orm ation pathways among hum a n populations or across speci es have practical implications for risk
assessm ent because d i fferent sp ecies of the same m e tals often have very different tox i cities. Because of this, the risk assesso r sho u ld fully exp l ore av ailable data on m etal m etabolism

. For
example, As can be m e tabolized to organic form s that are less toxic thanthe inorganic As to which an individual is initially exposed.

4.3.4.4. Excretion Risk assessors should be aware of the num b er of qualitative differences in m e tal excretion as compared to orga nic compounds. These include the greater likelihood for excretion in urine and the propensity for m e tals to be excr eted via hair and nails. In addition to reducing the target do se, risk assessors can u s e these excretion routes to develop biological m arkers of exposure for m a ny m e tals. Although metals share m a ny si milar characteris tics, their excretion kinetics can vary dram atically , prim arily b ecause of differences re g arding sequ estration in b one or bind ing to

4-21

### specific p r oteins. This reduces th e rate of excr etion (or generates a bipha sic excretioncurve with two or m ore distinct excretion half-lives). Risk assessors should assess each m e tal individually to inco rporat e exc r etion rates in th e c alculations of target dose and subsequent hazard
assessm ent.

4.3.4.5. Kin etic Mode ling Physiologically based pharm acokinetic/ pharm acodynami c (PBPK/PBPD) m odeling of
m etals entails the m athem atical description and modeling of their absorption, distribution,
m e tabolism , and excre t ion (ADME). A typical P B PK model consists of multiple co mpartm ents representing tissues or tissue gr oups that are linked by blood flow

. PBPD m odels describe the relationship between target tissue dose and health endpoints or targ et tissue effects. Com bined use of PBPK and PBPD m odels provides understanding of the complex relationships between exposure and target organ effects. Risk assessors m a y find that these m ode ls are valuable risk
assessm ent tools for purposes of interspecies, hi gh-dose/low-dose, route- to-route, and exposure scenario extrapolations (Kris hnan and Andersen, 1994). A PBPK model allows the risk assessor to define the relationship between external exposure and an intern al m easure of a biologically effective or toxic dose in both experim e ntal anim als and hum ans, thus increasing the precision of extrapolating effects thresholds to hum a ns. Use of PBPK models canaccount for nonlinear uptake, m e tabolism , and clearance; toxicity asso ciated with products of metabolism rather than the parent chemical only; and tissue interacti ons. The underlying assump tion of PBPK is tissue dose equivalence, i.e., that he alth effects are cau s ed by the to xic form (s) of the chemi cal
m easured at the biological target (Krishnan and Andersen, 1994). PBPK mode ls historically havebeen developed and used for risk assessm e nt m a inly with volatile organic compounds (VOCs) (e.g., m e thyl ene chloride) (Andersenet al., 1987), but PBPK mode ls have recently beenapplied to some m e tals (Wh ite et al., 1998; Clarke, 1995). There are cu rrently three m a in Pb risk assessm ent PBPK m o dels. The O’Flaherty model (O’Flaherty, 1995) is a P B PK model for children and adults. It includes the m ove m e nt of Pb from exposure m e dia (i.e., intake via ingestion or inhalation) to the lungs and gastrointestinal tract and subsequent exchanges betw een blood plasm a , liver, kidneys, and richly and poorly perfused tissues; and Metal Toxici t y 

### excretion from liver and/or kidney. The Leggett model The organ o rtissu e in whi c h metal toxic i ty occ u rs may differ from 

### (Leggett, 1993) allows risk a ssesso rs to simu late lif etim e the orga n or t i ssue(s) in wh ich the 

### exposures and can be used to predict blood Pb metal accum u lates a nd m a y be affected by the metal’s ki ne tics. 

### concentrations in both children and adults. The EPA has Effects at the portal of entrydo not 

### perform e d a comp arison of these adult Pb risk assessm ent depe nd on bi oavailability. Both the exposure rout e and the form of the 

### models metal can aff e ct a metal’s 

### ( http://www.epa.gov/superfund/lead/products/adultrevie carcin oge nic potential an d its non can c e r ef f e ct s.


### 4-22

### w.pdf ) . EPA developed the Integrated Exposure Up take 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 co rrective action sites (O SWER Directive, 1998; http://www.epa.gov/superfund/l ead/products/oswer98.pdf). Risk assessors can review differences in ki netic behavior between m e tals and VOCs in O’Flaherty (1998). In brief, these include: (1) oral bioavailab i lity, (2) i nhalation bioav a ilab i lity, (3) cellular uptake, (4) nutritionally essentia l and nonessential m etal in teractions, (5) proteinbinding behavior and function, (6 ) incorporation into bone or hair, (7) m e tabolism , and (8) excretion. Moreover, risk assesso rs should keep in mi nd that m a ny of the processes controlling the dispos ition of m etals are in trinsic ally c ap acity -limited and can resu lt in extended residence tim es. Risk assesso rs sh ould use mu ltip le lines o f evidence to understan dthe kinetics and, therefore, th e hazard of m etal sequestration and elimination. The m ajor challeng e faced by the risk assessor when using PBTK m ode ls for m e tals is to balance the comp lexity of the biology with the data available to param eterize the m odel. Estim ation of m any param eters from the sam e data or insufficient data (over-par am eterization) le ads to gre ate r uncertain ty in m odel predictions and limits th e utility of the m odel f o r regulat ory p u rposes.
4.3.5. Metal Toxicity Diversity in observed toxicities of differe nt m etals lik ely ref l ects the variety of biochemi cal m echanisms by which they exert th eir effects and variability in the i r to xicokine tic properties. At least five m e ta ls are known carcinogens, and seve ral other effects of metals are also well docum e nted, including effects on the neurological, cardiovascular, hem a tological, gastrointestinal, musculoskeletal, immunological, and epiderm a l system s .
4.3.5.1. Noncancer Effects of Metals Metals and m e tal comp ounds have very dive rse toxicological profiles. For risk
assessm ent purposes, selected critical effects serve as the b a s i s for deriv i ng thresho l dor benchm ark toxicity values (e.g., RfDs) and are defi ned as “the first adve rse effect, or its known precurso rth at occu rs to the m o st sensitiv e spec ie s as the dos e rate of an a g ent inc r eas es” (U.S. EPA, 2005). Both the mechanism of toxicity and the critical eff ect m ay vary with the form of the m e tal. Additionally, short-term exposures m a y produce target organ effects very different from those produced by a similar dose over a long er period of tim e. Shor t-term , high-level exposure by ingestion may give rise to well-recognized acute toxicity syndrom e s, usually involving the gastrointestinal tract initially and possibly, s econdarily involving the renal, cardiovascular, nervous, or hem a topoetic system s. Survivors of acute high-dose As ingestion usually expe rienc e mu ltiple organ ef fects, som eti m es with lo ng-term sequelae. Long -term , lowd ose exposu r es from ingestion of m e tals in food and water generally cause an accumu lation in

4-2 3

### target organs over tim e . Such exposures caninvolve any organ system but do not usually produce overt gastrointestinal symp tom s. For exam ple, low-level, long-term exposure to Cd in food—som e tim es com b ined with inh alation expo sure from cigarette sm oking—will cause Cd to
accumulate in target organs (e.g., kid ney) but wi ll not produce any obvious clin ical effects until “excess” capacity is dimi nished to a point wher e the norm a l function is lost (e.g., onset of renal disease and/or osteopo ro sis late r in lif e). In addition to considering systemic effects, the risk assessor should also examine portalof-entry effects. Unlik e systemic effects, which m a y be route-inde pendent, portal-of-entry effects are not observed following other routes of exposure. For example, derm al irritation, sensitization, and allergic responses from m e tals can occur without absorption and systemic responses.
4.3.5.2. Carcinogenic Effects o f Metals At least five transition metals—As, Cd, Cr(VI), Be, and Nil— are ac ceptedas hum an carcinogens in one form or another or in part icular routes of exposure (IARC, 2004b; NTP, 2002) and inorganic Pb compounds are consideredprobable hum a n carcinogens by EPA’s IRIS program , while IARC (2004a) has conc luded that there is limited ev idence of carcinogenicity to hum ans (see: http://www.epa.gov/iris/subst/0277.htm #carc and http://m onographs.iarc.f r /htdocs/announcem ents/vol87.htm ). Other m etals have mi xed evidence regarding potential carci nogenicity. Therefore, risk assesso rs should pay careful attention to approaches for cancer ris k assessm ent as app lie d to m e tals. Several gu idan ce docum ents are available for use by the risk assessor in developi ng or interpreting cancer risk assessm e nts (e.g., U.S. EPA, 2005), as are international efforts that provide guidance on assessing hum an relevance of

tum o rs identif ied in a n im als (Cohen et al., 2004). Nickel and Pb com pounds and Cr and Cr compounds ar e well-established contact allergens. O ther m e tals that have been cite d as contact allergens in clude Cu (WHO/IPCS, 1998), Co salts (AIHA, 2003), organom ercurials (A IH A, 2003), Be (IPCS, 1990b), palladium (Pd) (Kim ber and Basketter, 1996), a nd gold (Au) (Kimber and Basketter, 1996). Although there is
som e conne ction between skin and respiratory se nsitization, it does not follow exact rules, and the derm al mode is a much m ore common reaction to m e tals. A key consideration for the risk assessor in canc er risk assessm e nt is the determinatio n of the MOA of carcinogenesis, a general description of how the chemical causes cancer. The MOA determines hum a n relevance of observed anim al tum ors, any route-specific differences (e.g., carcinogenic at the portal of entry via the inhalation route, bu t not carcinogenic via the oral route), and the approach used for extrapolation from experimental doses in anim als to environm entally relevant hum a n doses. In partic ular, 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 som e , but not all, m e tals. For those for which the MOA is unknown, the risk assessor should refer to the Cancer Guidelin es (U.S. EPA, 2005) for guidance.

4.3.5.3. Issues Related to Evaluation of Toxicity Tests for Metals As with any hazard assessm e nt, risk assessors p r ef er a robust d ataset on tox i c respon ses of the m e tal(s) of concern for key endpoints (e.g. , irritation and sensitiz ation, systemic noncancer toxic ity, andgenotoxic i ty or tum o rigenicity). In m any cases, m etals will be well-s tud ied, and thus, hum an studies (epidemi ology, controlled clinical studies, or case reports) will be available to aid in hazard characterization. For m e tal compounds for which adequate hum an data are not available, th e risk assess or mu st rely on anim al toxicity studies. U.S. EPA has established guidelines for assessing the adequacy of a databa se for derivation of chronic hum an 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 m e tals as well as organic compou nds as long as the risk assessor considers the following metal-unique aspects of h azard determination:

• Adequate controls
. When a salt of a m e tal is ad mi nistered to the test an im als, the risk assessor should evaluate that a suitabl e control group was used, specifically that any potential for salt-induced toxic res ponses is appropriate ly assigned probable cause.

• Dosing solubility, ioniz ation, hydration, a nd speciation of metals adminis t ered in w a ter. Metal compounds m a y be in suspension or in solution and m a y be differentially hydrated depending on the concen tration in which they are prepared and the leng th of tim e the pre p aration stands, potentially resulting in different pharm acokinetic and toxic pr operties. W a ter pH and mi neral content also are relevant factors to be cons idered by the risk assessor.

• Trace element content of food and drinking w ater. Becau se of the well-known interaction o f m etals with essential tr ace elem ents, the trace elem ent content of anim al feed and drinking water or of vehi cles used for gavage or injection studies should be reported or controlled. Inconsis tent results across e xperim e nts could be due to this factor.

• Acute stres s in the exp eriment

. A component of acute stress in the experim e nt can induce hepatic m e tal-binding proteins (acute phase proteins) and alte r the toxicity of a given admi nistered dose. To achieve an adequate internal dose for the study of toxicity, anim al toxicologists often use bioavailable form s of m etals. For the initial charac terization of a toxicity syndrom e, it is not practical to simultaneously test all form s of a m e t a l that m a y be involved in hum an exposures. For example, Al research ers comm only use alumi num lactate, which is known to reliably provide elevated tissue concentrations in labo rato ry anim als, or aluminum m altolate, which 4-25

### provides a stable ion pool in wate r solution. However, a risk asse ssor is very unlikely to conduct
an assessm ent of Al in its lactate or m a ltolate fo rm . Thus, t he risk assessor should be aware that f

ailure to adjust the toxic ity data gene rated f r om water-so l uble m etal spec ies to the ap propriate, less soluble species of concern introduces uncertainty. There are both direct and indirect approaches to address the relativ e bioavailability of m e tals in the environm ent: (1) conduct new anim al toxicology studies using the m e tal for m e ncountered in the site assessm ent; (2) use adjunct scientific data to derive an adjustm e nt to the effec tive dose identified in the anim al study (e.g., data on the distribution of chemical for m s of the m etal in the environm ent or at a contaminate dsite ); o r(3 ) use a def au lt assump tio n that the metal in the e nvironm ental samp les is the sam e as that tested. The first approach is m ore scientifically sound but often is precluded by tim e and f i nancia l resou r ce limitatio ns; the third opti on generally is the m ost health-conservative. A fourth alternative to conducting site-specifi c assessm ents is for the risk assessor to estim ate bioavailability through so lubility studies or limited bioavailability studies of specific
sam p les from the site. For exam ple, arsenic bi o availability has been es tim ated f o r soils f r om various contaminated sites (Ng et al., 1998; Fr eem a n et al., 1995, 1993) and also through a series of solubility studies of soil from a site contaminated with mi ne 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 exposur e to the m e tal of interest. The risk assessor then uses these data as the input to the dose -response assessm ent to “characte rize 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-re sponse relationships, such as intensity and pattern of exposure, age, and lifestyle variables. Tr aditionally, separate approaches have been used for dose-response assessm en t for noncancer and cancer endpoints. For noncancer endpoints, the result of the dose-response assessment is an RfD for oral or derm al exposure or an RfC for inhalati on exposure. For cancer assessm e nt, the approach depends on the chemical’s MOA. Classically, th e result of the cancer asses sm ent is a measurem ent of the risk per unit dose, either as a slope factor or unit risk. A key consideration for the risk assessor when d eriv i ng m etal dose re spo n ses is to express the exposure potential and the toxic response as the sam e m e tal species. In general, for systemic effects of solub l e m e tal salts, the risk assessor shou ld express to xicity in term s of the dose of the m etal ion rather than the m etal salt. I n contras t , if the tox i city is re lated to a specif ic
compound, particularly for portal-of-e ntry effects, risk assessors s hould express it in term s of the
compound, r a ther than the ion. These differences should be considered when applying toxicity values in risk assessm ents. For example, in IR IS, Se is listed as “selenium and comp ounds,” and there are s ep arate assess m ents for “nickel, solubl e salts,” “nickel subsulfide,” “nickel refinery 4-26

### dust” (a mi xture), and “nicke l carbonyl” (a highly reactive co mpound that behaves dif f

erently from other nickel com p ounds). In som e situa tions, the toxicity of the anion needs to be considered. For exam ple, there are s e parate IRIS docum ents for Ag and silver cyanide. However, careful consid eration of th e chemical form has not been applied historically to all such docum ents, and the risk assessor shou ld car ef ully consider the applic ability of the chosen toxicity values to the chemical form s of interest, paying close atten t ion to solub i lity, bioavailability, and Consid era t ions in Risk Charac teri zati on 

### physical/chemi cal properties as well as Variability. Inter-i ndividua l biologi cal 

### available toxicity data. differen c e s e x ist within an animal o rhu man popul ation, or measu r em en t difference s exist 

### A related iss u e is whethe r the toxic owin g to method imp r e c isi on. 

### response is reported in term s of the added m etal Unce rtain t y. Data are u n a v ailable. Incertitude. Knowle dge a bout key 

### in the d i et o r the to tal m etal (i.e., whe ther the relation shi p s i s not availabl e. 

### am ount of tr ace elem ents inthe control diet in the anim al studies is included inthe dose calculations). Risk a ssessors should carefully review the supporting docum entation for toxi city values to determine prec isely what is being reported and to accou n t for any potential interaction s am ong dietary metals. Th e assessor also should consider whether the study provide d adequate lev e ls of trace e l em ents so that the ob served toxicity is not secondary to som e unrelated deficiency. Dose-respon se assessm e nts for som e m etals are based on data from hum a n occupational studies. While derivation of the dose-response value (e.g., the RfC, RfD, or cancer slope factor) will have in cluded som e steps to extrapolate the o ccupational study data to environm ental exposures (e.g., dosim etric and duration adjustm e nts), the form of th e chemical m ay m e rit consideration. For example, the occupational exposure will have involved a particular range of particle sizes, which m a y influence th e 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 shoul d not be below the general population RDA. This m eans that the risk assesso r sho u ld give c ar ef u l conside r ation to the appropriate size of uncertainty f actors, which is often made easier by the frequent (although not uniform ) availability of hum a n da ta for relatively large and diverse populations. T he only exception to the
comparison between the RfD and nutritional requ irem ents is that certain populations (e.g., pregnant wom e n) m a y have higher nutritional requirem e nts, while these levels could theoretically be toxi c to other populations. In such cases, the risk assessor should be careful to avoid log i ca l incons isten cies and to identif y the sensitive population on which the RfD is based.

4.4. RISK CHARACTERI Z ATION Risk Characterization is the final phase of the risk assessm ent process. It is the phase in which inform ation from hazard ch aracterization, dose-respon se asses sm e nt, and expo sure assessm ent is considered together to determine the actual likelihood of risk to exposed 4-27Risk 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 fr om 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, a nd the level of conf idence 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 wh ich 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 Characteri zation (U.S. EPA, 2000c) that identifies key goals and steps for a R i sk Char acterization. Each Risk Characterization should include three components: a qu alitative summ ar y of each section of th e risk assessm ent, a
num erical risk estim ate, and a de scription of uncertainties. S i n ce m e tal exposures often occur in the context of mi xtures (either mi xtures of m e tals of the sam e for m , mi xtures of different m e tal elem ents, or mi xtures with organics), risk assessors should consult additional Risk Characterization tools devel oped for mi xtures (U.S. EPA, 2000b, 1986). These guidelines specify that the characterization of risks from mi xtures of m e tals (and other compounds) be based prim arily on inform ation about the types of interactions that mi ght be present. Risk assessors should include a discussion in the Risk Characterizati on of the sources of variab ility and uncertain ty in th e risk assessm ent process. This is particularly important f o r m e tals risk assessm ents given all the components above. These are in add ition to the v ariab ility and uncertainties that are inhere nt in all risk assessm e nts (e.g., anim al-to-hum an toxicity extrapo l atio ns). Because inform ation, knowledge , and tools are lacking f o r m any of the m etalspecific uncertainties, risk assesso rs should be particularly dilige nt in docum enting whether these
m a y result in an over- or underestim ation of risk (i .e., result in a conservative risk estim a te or not). I tis likely thatsite -specif i c ris k assessm ents will h ave f ewer uncertainties th an regiona l- o r national-s cale assessm ents becaus e risk assesso rs have access to local data on key issu es such as specif i c m etal spec ies, r elative bio a vailab i lity, o r current m e tal levels. For national or regional
assessm ents, selection of ranges or specific num be rs for these values will depend on the degree

4-28

### of conservatism desired by the risk assessor a nd, therefore, should be clearly docum ented during the Risk Characterization phase.

4-29

### 5. AQUAT IC ECOLO G ICAL RI SK ASSES SMENT F O R METALS This chapter describ e s h o w to incorp orate the m etals risk assessm ent principles d escrib ed in Chapters 1 and 2 into ecological risk asse ssm ents involving aquatic-based receptors. Specifically, the followin g discuss i onfocuses on the relationship between each m e tal principle and components of the E P A’s Fram e w ork for Ecological Risk Assessm e nt (U.S. EPA, 1992a) and subsequent guidelines (U.S. EPA, 1998a). These comp onents include Problem Formulation, Characterization of Exposure, Ch aracterization of Effects, a nd Risk Characterization. The Problem For mulation phase consists of defini ng assessm ent endpoints, risk hypotheses, and a conceptual model to produce an analysis plan fo r the risk assessm ent. Chapter 2 d i scussed the consider atio n of the m etals princip l e s in th e Problem For mul ation pha se. In this chap ter consideration of the m e t a ls princi ples in the characterization of exposure, effects, and risk to aquatic organism s is discussed. Consisten twith the p r e v iously s t ate dsc ope and purpose of this Fram ework, not all aspects of th e ecolog i cal risk as sessm ent process are discussed. On ly those aspects of the aquatic ecological risk assessm ent proces s and associated t echnical issues with the greatest relevancy to the m e tals principles are incl uded. Also empha sized is how th e geographic scale (e.g., site specific, regional, national) and analytical sc ope (e.g., screening vs. de finitive analysis) of aquatic risk assessm ents affect the ex tent to wh ich the m e tals principles can be incorporated. Although these principles apply e qually to risk assessm ents invol ving terrestrial-based ecological recep tors, m a ny of the methods and tools that can be used to implem e nt these prin cip l es differ between the aquatic and terres trial environm ents. Thus, a separate discussion of how the m e tals principles apply to ecological risk assessm ent in the terrestri al environment is provided in Chapter 6.

5.1. METALS PRINCI PLES Metals have specific environm ental and biotic attributes that should be considered in all risk as sessm ents. Specifically, these attributes fa ll into the ri s k assessm ent parad i gm as f o llows:
Background levels Exposure Assessm ent Mixtures Exposure and Effects As sessm ent Essentiality Effects Assessm ent Form s of m e tals Exposur e and Effects As sessm ent Toxicokinetics/toxicodynamics Exposur e Assessm ent (bioavailability) and Effects Assessm ent (absorption, distribution, m e tabolism , and excretion [ADME] and toxic ity)

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 analys is 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 ecologi cal risk assessments is found in U.S. EPA (1998a, 1992a). Exposur e Profile The expo su re profile sh oul d descri be the exposure pathways fro m stre ssor so urce to the recepto r , the exposu r e intensity, its spatial and temporal distribution of co

occurre n ce, and the impa ct of variability and uncertainty on the exposure estimate s.



### 5.2.1. Background Lev e ls Background levels refers to those con centration s of m e tals that derive from natural as well as anthropogenic sources that are not the focus of the risk assessm ent. In aquatic ecosystem s, m e tal concentrations vary widely over space and tim e owing to differences in watershed geology, hydrology, anthropogenic and natu ral loads from “nontarget” sources, and other factors. Depending on the m a gnitude of the exposure associated with these factors, background m e tal concentrations can account fo r a sign ificant portion of total m e tal exposure. Furtherm ore, certain es s entia l m etals can bi oaccumulate to high levels in som e aquatic organism s (e.g., Zn in barnacles, Cu in cray fish) due to species-sp ecific physiological requirem e nts, regard less of source. Even som e nonessential m etals can naturally b i o accumu late to signif i can t leve ls as a result of mi mi cry of ess e ntial m e tals or sequestration and storage. Thus, the risk assessor needs knowledge of background concentrations in order to characterize exposure and to differentiate risk associated with m e tal sources already in the environm ent from risk as sociated with the m etal source s of concern in the assessm ent. En v i ronmental Backg rou nd in Expos ure 

### Depending on the design and context Ass essme nt 

### for the assessm ent, the risk assesso r needs to What a r e the environ menta l backgroun d con c e n tration s at the site(s) of intere st? Ho w do 

### address several questions and issues they vary over releva nt spa t ial and temp oral scale s ? 

### pertaining to background concentrations of What is the relation ship b e twee n 

### m etals. Obtaining reliab le es tim ates of environ menta l backgroun d and toxicol ogi cally relevant meta l con c entratio n s? 

### background concentrations can be Can n atural a nd anthropo g enic metal b e 

### challenging, particularly at larger spatia l disting u ished ? To what exte nt are ba ckground con c e n tration s 

### scales. R i sk assesso rs are cautioned agains t being extra p o l ated over sp ace a nd time ? 

### using background m e tals in soils as surrogates What level of confid en ce (u nce r tainty) ex ists in the estimate of environm e n tal backg rou nd 

### for sedim e nts due to differences in the c o nc en tr a t io ns ? 

### biogeochemi cal processes between these two environm ents. Meta l co ncentr ation s in

5-2

### sedim e nts can be impacted by sedim e nt physic ochemi cal composition and localized sedim e nt transport processes. Fractionation of sedim e nt co res, in com b ination with caref ul es tim ates of sedim e ntation rates, has been used to distingu ish pre- vs. post-industria lized contributions of
m e tals via atmospheric transport (e.g., m e rcury) (U.S. EPA, 1997b). Ri sk assessors may find this type of sedim e nt core analysis useful fo r differentiating n a tural levels versus levels associated w ith anthropogenic sources. Concentrations of m e tals in the water column vary over tim e and are highly responsive to hydrological changes. In site-specific risk assessm e nts, the risk assessor m a y quantify background levels by m e asuring m e tal concentrations at sites upstream from the area of concern. National databases of metal concentrations in va rious aquatic m edia (sedim ents, water, biota) include the f o llowing:

• The EPA STOrage and RETrieval (S TORET) database (accessible at http://www.epa.gov/storet/);

• The National Sediment Quality Survey (NSQS) (accessible at http://www.epa.gov/waterscience/cs/nsidbase.htm l);

• Environmental Monitoring and Assessmen t Program (EMAP) (accessible at http://www.epa.gov/em ap/);

• The National Stream Quality Accounti ng Network (NASQAN) (accessible at http://water.usgs.gov/nasqan/), and

• The Hydraulic Benchmark Ne twork (HBN) (accessible at http://ny.cf.er.usgs.gov/hbn/).
The risk as s e ssor shou ldkeep in mi nd that with the exceptionof the HBN, these databases have not been established to dis tinguish background concentrations from local, anthropogenic sources of m e tal loadings (e.g., in dustrial discharges, stormwater runoff) and, thus, m a y reflect significant anthropogenic loadings of m e tals to environmental m e dia. The HBN was established to provide long -term m easurem ents of str eam flow a nd water quality in areas that are mi ni m a lly affected by hum a n ac tivities. While the HBN contains long-term
m easure m ents of a num ber of param eters that ar e known to affect the bioa vailability of m e tals (e.g., dissolved organic carbon [DOC ], inorganic ions such as Caand Mg, pH, conductivity), it does not contain inform a tion on m e tals of typica l regulatory concern, withthe exception of Al.

5.2.2. Forms of Metals The physical and chemical form s of m etals affec texposure, bioavailablity, and subsequent effects and are influenced by physic ochemi cal environm enta l conditions. Nationallevel assess m e nts involve a broad range of envir onm ental conditions ; th erefore, th e risk as sesso r
5-3

### should acco unt for different m e tal species in di f f

erent locations and water body types. As
assessm ents transition from national, to regional, and to local, the a ssessor should incorporate site-specific sedim e nt and water quality pa ram eters that influence m etal speciation,
complexation, and sorption onto biological surfa ces (e.g., pH , organic carbon, inorganic ligands, Ca, Mg, sulfide). Speciation m odels (e.g., MINE QL) com bined with biotic ligand m odels offer a fra m ework for addressing the differential occurren ce and toxicity of various m e tal form s. Risk assessors should be aware of the difficulty in applying this approach to assessm ents involving large regional or national scales because of the variability in model param e ter values (including covariance among parameters) that occurs across locations. Infor m ation about the range of the input parameters can be derive d from available databases (s ee Section 5.2.1), and the risk assessor can decide what value to use (e .g., mi nimum , m axi mum , m ean) depending on the degree of conservatism desired in the assessm en t. The risk assessor should include this inf o rm ation in the Risk Characte r iz ation as p art of the overall discus sion of assump tions and uncertainties in the asses sm e nt. Risk assesso rs can directly as sess the m e tal form s for sitespecific assessm ents or estim ate what these would be based on sedim e nt/water param e ters (see Section 3.1.3, Environmental Chemistry).

5.2.3. Exposure Pathway Analysis For aquatic organism s, potenti al routes of exposure to
m etals inclu d e absorptio n across (or in som e cases adsorption to) respiratory organs, derm al absorption, sediment ingestion, and food ingestion. Quantifying exposure and uptake by the respitory route is a particularchallenge to aquatic risk assessors because of the differi ng types of respiratory organs
am ong aqua tic species, the dynami c nature of the respiratory process in water, and the intim ate co ntact b etween the recep tor andm etals diss olved in waters. Furth er comp licating the issue, som e respiratory organs can also be involved in locom otion, excretion, ion regulation and the capture, sorting, and ingestion of food. Similarly, ri sk assessors m a y find the ingestio n route d i fficult to defin e for aquati c recep tors because of th e divers ity of feeding modes and food sources, such as sedim e nts, susp ended solids, mi croflora, anim al tissues, and plant tissu es . The use of stable iso t o p e techn i que s has contrib u ted gre atly to evalu atin g the ro le of diet in co ntaminant accumu lation (including m e tals) by precis ely defining troph ic interactions (e.g., Kidd et al., 1995; Jarm an et al., 1996). Th e absorption route can involve uptake across a phytoplankton cell m e mbrane, amphibian skin, arthropod exoskeleton, the egg m e m b rane, or the integum ent of an infaunal clam or annelid. Despite the com p lexities associated with quantifying exposu r e of aquatic anim als to
m etals from multiple rou t es of uptake, risk as se ss ors can find a significant am ount of i nfor m ation 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; Bjerrega ard et al., 1985). Applications of one-compart m e nt biokinetic models using laboratory-based m easurem ents of key m odel param eters (assimilation efficiency,
m e tal uptake rates from water and food elimination rates) have been extended to field situations for populations of a diverse array of aquatic spec ies, including freshwaterand m arine bivalves, various crustacean s suchas copepods, amphipods , and crab, aquatic insects, and fish (e.g., Luom a and Rainbow, 2005; Stewart and Fisher, 2003; Griscom et al., 2002; Baines et al., 2002; Fisher et al., 2000, 1996; Roditi et al. , 2000; W a ng et al., 199 6; Luom a et al., 1992). S itespecific m odel pred ictio n s for m etal concentratio ns in anim al tissues are strikingly close to independent field m easurem ents for diverse wate r bodies, suggesting that it is possible for risk assessors to account for the m a jor processes gove rning contaminant concen trations in aquatic anim als and that laboratory-deri ved kinetic parameters are app licable to natural conditions (Luom a and Rainbow, 2005). Thus, these m odels pr ovide tools for risk a ssessors to use when addressing m e tal exposure and uptake, and they can be used to determine the relative importance of different routes of exposure (Lan drum et al., 1992; W a ng e tal., 1996). Risk assessors should always consider tempor al aspects of exposure, particularly in aquatic system s that respond to frequent shif ts in hydrology. Rapid speciation and phase changes associated with changes in pH/Eh m ake te mporal issues particularly germ ane to m e tals. Fluctuating or pulsed exposures oc cur in situations such as rapi d changes in pH/Eh ass ociated with photosynthesis and respiration, hypolimnetic di scharge from stratified reservoirs, biocide (e.g., copper sulfate) spraying, ingestion of prey item s with seasona lly high m e tal concentrations, surface waters receiv ing wastewater treatm e nt pl ant effluent, urban sto r m water, snow m e lt, and acid precipitation runoff. Transien t m etal concentrations m a y be orders of m agnitude higher than typical or average concentrations but lastfor only a few hours. These episodic exposure scenarios have been poorly characterized for m e ta ls (Butcher et al., 2006). Any risk assessm ent for m e tals should clearly state al l assumptions about duration of exposure and what uncertainties are added to the risk m o del as a consequence.

5.2.4. Fate and Transport of Metals Risk assessors routinely use transport and fate computati onal models to describe and quantify exposure pathw a ys. Models also are usef ul in situ ations where risk assesso rs need an estim ate of future exposure levels that are expected to result f r om the i mplem entation of som e permitting a ction o r rem ediation m e asures atlo c al, reg i onal, or national s cales. Numerous models are available for use; m o st are based on the sam e fundam ental principles. Metals are ubiquitous in the environm ent and within ea ch m edia comp artm ent they are present in association with air, water (freely dissolved m e ta l or as organic and inorganic m e tal complexes), and particles (sorbed, precipitate d, or incorporated within a mi ne ral ph ase). The risk assessor can find a detailed discussion of the fate and tr ansport of m e tals in Section 3.2. No single, 5-5

### curren tly av ailab l e m o del inc l udes a ll th e desirabl e m e tal-sp ecific featu r es for aquatic system s. Discussions of the fa mily of available aquatic tr ansport and f a te m odels, as well as a num b er of chemical equilib rium models, m a y be found in Paquin et al. (2003).

5.2.5. Toxicokinetics a nd Toxicodynamics (B ioavailability and Bioaccumulation Issues) 5.2.5.1. Aqueous Phase In the d i ssolved phase, m etals can e x ist as f r e e ions as well as in a variety of complexed for m s. These for m s, or species, are of key im port a nce in understanding bioavailability, and the hazard and risk assessm ents of waterborne m e tals are comp licated by the f act that m etal species differ in their toxicological properties. For m a ny m etals in aq uatic system s, it is the f r ee ionic for m that is most responsible for toxicity. For example, Cu 2+ 

### has been directly linked to toxicity in fish and invertebrates while Cu comple xed by dissolved organic m a tter does not induce toxicity to th e sam e degree (Ma et al. , 1999; Erickson et al., 1 996) because of its redu ced availability f or uptake by the organism. Howe ver, the risk as sessor shou ld be aware that although toxicity of m e tals bound to DOC is redu ced, it is not eliminated entirely and can contri bute to the to tal m etal loading to an ecosyst em and subs equent toxic effects (McGeer et al., 2002; Erickson et al., 1996). On the other hand, there are cases wh ere nontoxic m e tal species are bioavailable and taken up by the organism but cause no adverse response (e.g., Ag-Cl comp lexes in rainbow trout) (McGeer and W ood, 1998). Risk assessors should recogni ze that the presence of m e tal within an organism cannot always be used as a surrogate for toxic response. Risk assesso rs can choose am ong a variety of m e t hods to acco unt for relative bioavailability of m e tals in a quatic system s, including hardness adjus t m e nts, water-ef fect ratio (WER), Free Ion Activity Model (F IAM), and aquatic Biotic Liga nd Models (BLMs) (Paquin et al., 2002a). Each m e thod contains strengths a nd limitations and m a y not be am enable to all types of assessm ents (e.g., ranking /classification, national, and site -specific assessm e n ts). For
example, adjustm e nt of a queous m e tal concentr ations for diffe rences in water hardness was
am ong the fi rst computational m e thods to accoun t for bioavailability d i fferences b e tw een the laboratory and the field when applying EPA wat e r quality criteria. Although these adjustm e nts are relatively easy to ap ply, they req u ire empirica l data to define the toxicity-water hardness relationship. Thus, they are m o re amenable to site-specific ri sk assessm ents, although even in those cases they do not account for other water qu ality factors that aff ect bioavailability (e.g., DOC, pH). The water hardness approach has be en applied at the national level through a statem ent of water quality criteria as hardness-based equations ratherthan as single values. Risk assessors can choose to use ranges, m eans, or median values when conducting large-scale (e.g., regional) assessm ents, but they will need to acknowledge this uncertainty during the risk characterization phase.

5-6

### Bioavailability adjustm e nts using WERs inco rpo r ate the com b ined effects of all water quality param eters prese n t in site water on b i oav ailab i lity of m etals of interest re lativ e to what
was m easured in laboratory tests (U.S. EPA, 199 4c). Thus, compared to hardness adjustm e nts, WERs encompass a broader array of water quality factors that can impact bioavailability. However, W E Rs are re lative l y resou r ce intensive (re quiring toxicity tes tin g), are applicable only on a site-specific basis, and are not e asily adjus t e dto accoun t f o r tem poral or spatia l v ariab ility. This relationship between sp eciation and bioavailability al so has been explained through the free ion activity m odel (FIAM, C a mpbell, 1995)

. This model produces speciation profiles of
a m e tal in an aquatic system and provides insigh t into the r e lative bioav ailab i lities o f the different form s of m etal as well as the im portance of complexation. Models available for the calculation of m e tal speciation in natural waters are review ed in Section 3.1.5, Sediment Che mistry, and include MINEQL (Schecherand McAvoy, 1994; W e st all et al., 1976), MINTEQA2 (Brown and Allison, 19 87), CHES S (Santore an d Driscoll, 1 995), WHAM (Tipping, 1994), and PHREEQ (Parkhurst et al., 1980). The risk assessor should review Paquin et al. (2003) for a m o re in-depthunderstanding of these m ode ls and how to select am ong them for particular places and types of assessm ents. The BLM approach successfully combines th e influences of speciation (e.g., free m e tal ion, DOC complexation ) and cationic com p etition (e.g., K + 

### , Na + 

### , Ca +2 

### , Mg +2 

### ) on m etal toxicity in fish (De Schamphelaere and Janssen, 2004, 2002; De Schamphelaere et al., 2004, 2003, 2002; Heijerick et al., 2002a, b; Di To ro et al., 2001; Santore et al., 2001; McGeer et al., 2000). The model can be used to distinguish, at least conceptually, m e tals thatwill bioconcentrate at the site of toxicity (e.g., gill or other biotic ligand) from t h e total m e tal pools in an organism and the bioavailable m e tal pool in the exposure m e dia. The m odel also can be applied to algae (De Schamphelaere et al., 2003; Heijerick et al., 2002) and Daphnia (De Schamphelaere and Janssen, 2004, 2002; De Scha mp helaere et al., 2004, 2002). The BLM recently has been incorporated into dr af t re visions to E P A’s nation al wate r qua lity criteria f o r Cu (it h as been used in risk
assessm ents at a range of geographic scales), and it is b eing applied a s an alte rnative to the W E R approaches for setting site-specific discharge ob jectives. The BLM has the potential to address spatial and temporal factors thataffect bioavailability, providedthat the variab ility in water quality param e ters used as inputs to the m ode l can be quantified or predicted (e.g., pH, DOC, K + 

### , Na + 

### , Ca +2 

### , Mg +2 

### ). However, the risk assessor shoul d be aware of the many limitations in applying the BLM (or any Free Ion model). Fo r example, the developm ent of the BLM has focused prim arily on bioavailability and acute e xposures. Work has begun to extend the BLM to chronic toxicity for som e organism /m etal com b inations (e.g., De Scha mphelaere and Janssen, 2004; Paquin et al., 2002a, b), and further development is expected. Also, the BLM is currently based on m e tal uptake through the dissolved phase; thus, additional research is needed to address
m e tal uptake and toxicity via the diet.

5-7

### 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 sediment s, sulfides provide the primary binding phase for many cationic metals. These metal sulfides ar e 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 relati vely nontoxic. The amount of reactive sulfide is quantified by measuring the amount of sulfide freedwhen 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. Wh en SEM-AVS < 0, sufficient sulfide exists to bind all SEM, and metal toxicity is not expect ed. When SEM-AVS > 0, metal is present beyond the binding capacity of su lfide, and toxicity may occur if th ere 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 beenexplored 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 additi on, several studies have shown some degree of metal accumulation in organisms exposed to sedime nts where sulfide is in excess and metals are thought to be nonbioavailable, or at least nontoxi c (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 as sessors 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 dete rmine the bioavailable concentrations of Sediment-As s ociated Me tals Whe n the mo lar co ncentrations of acid -volatile sulfide in se di ment excee d the a m ount of sim u ltaneo usly extracted m e tal, the metals are expecte d to be asso ciated with the solid p h a s e a nd not to be toxic.



### sedim e nt-bound m e tals include m e tal concentrati ons in the chemical extracts (Fan and W a ng, 5-8

### 2001; Babukutty and C hacko, 1995; Tessier et al., 1984), acid extracts (Langston, 1980; Luom a and Bryan, 1978), or biomi m etic extracts (W est on and Maruya, 2002; Mayer et al., 2001; Chen and Mayer, 1998). However, no consensus yet exis ts on their best use fo r different types of
m e tals or metalloids. S e veral other m e thods have been proposed. Based on the premise that iron oxides in oxic sedim e nts lower m e tal bioavailabil ity, Fe in a 1 N HCl sedim e nt extract has been used to norm alize m etal exposure concentr ations (Luom a and Bryan, 1978). Increasing concentrations of organic carbon can decrease me tal bioavailability (Crecelus et al., 1982), so
norm a lization of sedim e nt m e tal concentrations to organic carbon content has been conducted in other c ases. The m o re readily extr a cted m eta ls from sequential chemical extraction schem e s tend to be the m ost bioavailable (Young and Harv ey, 1991; Tessier et al., 1984) and have been used to es tim ate bioavailab l e m etal.

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 pot entially toxic effects (eithe directly to the aquatic organi sm or its predator). The main reason for this ambiguity is that the bioavailability of dietary metals varies wide ly across organisms and exposur e conditions and standardized approaches are not available fo r 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 bereleased from the ingested particle into the gastrointestinal fluids of the animal (C hen 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). Assi milation efficiency (i.e., the net amount of metal retained in tissues relative to the amount ingested from f ood) 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 toxi city 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 depend ing on the metal, its form and distribution in prey, species digestive physiology (e.g., gut re sidence time), environmental conditions (e.g., temperature), food quality, food ingestion rate, and metal concentration in the diet. Thus, risk Dietb orne M e tal Expos ur e Risk a s sesso r s shoul d co n s ide r dietborne met a l exposure i n two contexts: (1 ) dietborne exp osu re leadin g to accumul a tion a nd exposure to h igher level s in the food chai n (e.g., human s, wildlif e) and (2) dietborne exp o su re lea d ing to direct 

### r effects on ex posed organi sm s.



### 5-9

### assessors should consider likely ra ng es of assimilation efficiencies for a p articular m etal-an im al
com b ination when evaluating m e tal bioavailability from the diet. A number of reviews have summarized current knowledge of assimilation ef fi ciencies of ingested m e tals am ong different aquatic anim al species (W ang and Fish er, 1999; Fisher and Reinfelder, 1995). The distribution and form of m e tals in diet ary organism s is of critical importanc e f o r understanding the bioavailability of dietary m e tals and trophic transfer pote ntial. For example,
m e tals in the cytosolic fraction of phytoplankt on and “soft parts” of zooplankton have been shown to correlate well with bioaccumulated me tal in their predators (e.g., herbivores and plantivorous fish, respectively) (W ang and Fi sher, 1996; Reinfelder and Fisher, 1994a, b). Metals sorbed to the cell wall of phytoplankton and the exoskele ton of zooplankton w e re poorly assimilated by consum e r s. The bioavailability o f m e tal-enriched granu l es in prey (a detoxification and storage m echanism exhibited by som e organism s) has been shown to be negligible or substantially reduced w hen cons umed by certain predators (e.g., W a llace et al., 2003, 1998; Wang, 2002; W a ng and Fisher, 1999; Nott and N i colaidou, 1990). However, the risk assessor should be aware that the bioavailability of m e ta l-enriched granules in prey can vary
am ong metals and with type of granule (W ang, 2002; Mason and Jenkins, 1995; Nott and Nicolaidou, 1990) and may also depend som e what on digestive physiology of the predator (e.g., gut pH, retention tim e ). As a result of thes e findings, fractionating body bur dens of m e tals (e.g., cytosolic m etal vs. m etal granules) has been s uggested as a better m eans of identifying the bioavailable fraction of dietbo rne metals (Seebaugh and W a llace, 2004 ; Wallace et al., 2003; Fisher and R e infelder, 1995; Reinfelder and Fisher, 1994). Although such techniques show promise for operationally defining the extent to which dietary m e tals m a y be bioavailable for trophic transfer in aquatic food webs, risk assessors should unders tand that broad-scale application of these techniques to m etals risk asse ssm ents is presently limited by the re latively
sm all num b er of m e tals and predat or-prey relationships evaluated. Despite the uncertainties associated with bi oavailability and trophic transf er of dietary
m e tals, the use of whole-body inorganic m e tal con centrations in prey species m a y have som e utility to risk assessors for conservatively screening for exposure and potential risks to
consum ers (i.e., in cases where whole-body residues are below dietary toxic th resholds ). For m o re def i nitiv e as sessm ents, further res ear ch is need edto Trophic Tran sfer 

### quantify the bioavailability and effects of inorganic Trophic tr an sfer i s the tra n sfer of a 

### dietary m etals, with th e e x ception of certain chemi c al fro m prey sp eci es to a predat or spe c ie s via di etary exposure. 

### organom etallics (e.g., methyl m e rcury) and Biomagnific a tion is a typ e of trophic transfe r where chemi c al co nce ntration s 

### m e talloids (e.g., Se) where dietary toxicity has been increa se in o r gani sms from a lower 

### well es tablis hed. trophi c level to a highe r tro phic level within the sa me food we b. Biodilution repre s e n ts a d e crea se in orga nism co n c entration wit h increa sing trophi c level. 

### 5-10

### 5.2.5.4. Bioaccumulation and Trophic Transfer Assessing and predicting the bi oa cc umulation o f m etals in a quatic ecosy stem s is a
component of m a ny Agency regulatory and nonregulatory activities (e.g., chemical ranking/classification, derivati on of national water quality cr iteria, Superf und site risk
assessm ents; see Section 2.3.1 for definitions and a conceptual m odel related to bioaccumu lation). Interest in m etals bioaccumu lation originates from concerns regard ing the direct impact of m etals o n organism s accumu latin g the m etal and indirect impacts on their
consum ers (i.e., trophic transfer). Un like cert ain persisten tan d bioaccumu lative organic
compounds, which tend to biom agnify in aquatic food webs (e.g., DDT/DDE, PCBs, PCDD/PCDF), inorganic m e tal compounds rarely bi om agnify across three or m o re trophic levels (McGeer et al., 2004; Suedel et al., 1994); however, certain organom etallics can biom agnify in aquatic food chains. Som e me tals (e.g., lead) tend to biodi lute in aquatic food webs. Risk assessors should not interp ret lack of biom agnification as lack of exposure or concern via trophic transfer (see text box). Ev en in the absence of biom agni fication, aquatic organism s can bioaccumu late relatively large am ounts of m etals and becom e a significant source of dietary
m e tal to their predators (R einfelder et al., 1998). For m a ny nonionic organic chemicals, ri sk assessors can derive first-order approxim a tions of bioaccumulation potential fr o minfor m ation on chemi cal prop erties and organism attributes (e.g., K ow 

### and lipid content) and their use as inputs to simp lified, fugacitybased m odels (e.g., Gobas et al., 1993). For m e tals, analogous m e thods to predict bioaccumu lation bas ed o n sim p le chemical prope rties are no t a v ailab l e or are not widely validated. T he lack of analogous models for m e tals is likely due, at leas t in part, to the high degree of specificity exhibited by the m echanisms and processes underlying m e tals bioaccumu lation (e.g., sp eciation, exp o sure cond itions, and organism phys iology) (see McGeer et al., 2004; Rainbow, 2002; Mason a nd Jenkins, 1995). As a result, risk assessors currently are limited to using an empirical approach for a ssessing and predicting m e tals bioaccumulation. Typically, this requires direct m easurem ent of me tal concentrations in the organism or experim entally-d etermi ned param eters for use as input to b i oaccumu lation m odels (e.g ., gill uptake r ate, elimination rate, assimi lation efficiency). Aquatic ecological risk asse ssors commonly use bioconcentr ation factors (BCFs) and bioaccumulation facto r s (BAFs) to quantify ch emical accumu l ation in tissu e relativ e to concentratio n in water. BCFs and BAFs are determined as the ratio of the chemical concentration in tissue to its co ncentration in water (using the st eady-state m e thod) or as a ratio of uptake rate (k u 

### ) and elimination rate (ke 

### ) constants (using the kine tic method). Measurem ent of BCFs or BAFs usually is conducted for condi tions that approxim a te steady-state (i.e., where
accumu latio n rem ains relatively con s tant due to chemical uptake being of fset by its elimination by the organism ). It is assum e d that the greater the BAF or BCF, the greater a chemical’s bioaccumu lation or b i oco n centration potential. 5-11

### Risk assessors should recognize th at considerable uncertainty can be associated with the applic ation of liter a tur e - d erived BC Fs and BAFs f o r assess ing the ri sks of m etals, as variab ility in BCFs and BAFs for metals is known to be high (e.g., 50-fold or higher within a metal). Most of

this uncertain ty resu lts f r om bioavailab i lity differences between the studies from which the BCF or BAF is m easured and the site(s) to whic h it is b eing applie d (e.g., water quality characteristics, m e tal speciation, exposure pathw a ys). Other sources of uncertainty that risk assesso rs should consider in the broad application of BCF/BAF data are rooted inthe complex m e c hanism s of m e tal toxicokinetics (uptake,
m e tabolism , distribution, elimination). Fo r example, unlike hydrophobic, nonionic organic chemicals where uptake across biological m e m b rane s generally occurs via passive diffusion, the uptake of metals is believed to involve a num be r of specific transport m e chanism s. Som e of these transport m echanism s involve binding withm e m b rane carrier proteins, transport through hydrophilic m e mbrane channels, and e ndocytosis. Passive dif f usion is thought to be reserved for certain lipid soluble form s of m e tals, such as alky-m etal comp ounds and neutral, inorganically
complexed m e tal species (e.g., HgCl 2 0 

### ). The implication of th ese spec if ic transpo rt m echanism s is that m etal bioaccumu lation can in volve satura ble uptak e ki netics, suchthat BCFs and BAFs depend on exposure concentration. The existence of saturable uptake m echanism s, the presence of significant am ounts of stored m e tal in organi sm s, and the ability of som e organism s to regulate bio accumu lated m etal with in certain ran g es are all th ought to be responsible for the inverse relationship that has b een frequently reported betwee n BCFs/BAFs and m e tal exposure concentrations (McGeer et al., 2003; Borgm a n et al., 2004). In these cases, higher BCFs or BAFs are associated with lower exp osure concentr ations an d also can b e associated with lower tissue concentrations within a given BCF or BAF study. This is counter to the implicit
assump tion that higher BCFs or BAFs indicate h i gher m e tal hazard. As a result o f the aforementioned un certainties , risk assessors must be careful in broadscale application of BCF and BAF data for m e ta ls. Spec if ically, the current science does not support the use of a single , generic threshold BCF or BAF va lue as an indicator of m e t a l hazard. Similarly for national risk assessm ents, use of a single BCF or BAF value holds little utility due to high uncertainty that results from differen ces in bioavailability, exposure conditions, and species -specific factors that influence m e tal bioaccumulation by aquatic organism s. When extrapo l atio n across s ite s is nece ssary and limite d data pr eve n t applicatio n of alternat ive approaches (discussed below), uncertainty in the use of BCFs and BAFs can be reduced by expressing them as a function of m e dia chemistry (i.e., to address bioavailable m e tal), exposure concentrations (i.e., to address concentration dependency issues), and limiting extrapolations to within a particular spec ie s or closely related species. The us e of BCFs and BAFs for metals
assessm ents appears to h a ve m o st value for site-s pecific app l ication s, when appropriate
m easurem ents are taken f rom the site (s) of inte rest and extrapolation of BCF/BAF val ues across differing exposure conditions a nd species is mi nimi zed. 5-12

### Risk assessors should be aware of seve ral alternatives for assessing m e tals bioaccumu lation th at addr ess som e of the concern s lis ted abov e. One of these is to dev elop regression relationships between tissue and exposure concentrations. Such regression relationships have been used to characterize bi oaccumulation of m e tals by soil organism s (U.S. EPA, 2003c; Sa mple et al., 1999), but they have not yet been comp iled for aquatic system s. The advantage of this technique is that it addre sses the dependency of BCF or BAF on exposure concentration. However, it does no t explic itly ad just f o r bio a vailab i lity d i f f

erences th at occu r across s ites. Another alternative is to use a kinetic-based m o del for describing bioaccumulation (Luom a and Rainbow, 2005; W a ng and Zauke, 2004; Kahle and Zauke, 2003; Chang and Reinfelder, 2002; Reinfelder etal., 1998). These m odels can improve predictions of m e tal bioaccumu lation in aquatic organ i sm s because th ey incorporate different exposure rou t es (e.g., water vs. diet) and the dy nami c nature of m etal bioaccumu lation processes

. For example, Luom a and Rainbow (2005) reviewed the DYNBAM model (a single-c ompartm e nt, kinetic-based bioaccumu lation m odel) and found it to accu rately predict m e tal b i oaccu mulation for a wide range of m e t a ls, organism s, and habitats based on data derived from 15 separate studies. Importantly, DYNBAM and similar such m odels require experim ental data m easured under environm entally-realistic conditions in order to de rive m odel param eters for each m etal-species
com b ination (e.g., uptake and elimination rates, a ssimilation efficiency, food ingestion rates). Comp ilations of such data on m odel input param e te rs are available for some species and m e tals (e.g., W a ng and Fisher, 1999, for aqua tic invertebrates). Clason et al. (2004) and Kahle and Zauke (2003 ) have developed two-co mpartm ent bioaccumu lation m odels for amphipod crustaceans that incorpo r ate b ackgro und m etal a nd saturatio n of uptake kinetics. Cu rrently, however, th ese m odels include only the disso lvedphase and do not accou n t for uptake from the diet. The bioaccu mulation m o dels describ e d above offer strong pro mise for i mproving bioaccumu lation predictions in a quatic risk assess m e nts for metals and sh ould be cons idered by risk assessors. However, risk assessors should be aware of their limitations. For example, they
curren tly do not accoun t f o r dif f

erential partition i ng and bioa vailab i lity o f m etal in or ganism s (see Figure 2-2). Empirical m e thods are being developed to pr edict m etal com p artm e n talization in tissues of aquatic organism s (e.g., Wallace an dLuom a, 2003; W allace et al., 2003 ), but these have not bee n incorpor ated into b i oa ccumu lation m odels. These m odels also do not e xplic itly address the impact of m e tal sp eciatio n on bioaccu mulation or link bio accu mulated m etal to toxic effects, although such models ar e under developm ent (Paquin et al., 2002b). Regardless of the type of bioaccumu lation m odel used, reductions in uncertainty in m e tals bioaccumu lation
assessm ents should be directed at ac hieving robust connections between the bioaccessib l e/bioav ailab le form (s) of m etals in v arious exposu r e m edia, their accumu lation,
m e tabolism , and distribution in tissues, and the form (s ) of m etals th at exe rt their toxicity directly to the organism or indire ctly to its co nsum ers. 5-13

### 5.3. CHARACTERI Z ATION OF EFFECTS 5.3.1. Essentiality Essentiality ref ers to the nutritiona l r equire m e nts of an organism for nor mal m e tabolic function. A key difference between m e tals and or ganic chemical contaminants is that som e m e tals are required either as m acronutrients (e.g., Fe, Ca, Mg) or mi cronutrients (e.g., C u, Zn, Ni) to m a intain a healthy organism

. Table 6-3 inSection 6.2.1 classifies the m e tals addressed in this Fram ework by their known essen tiality to plants and an im als. Consideratio n of essentia lity by the ri sk assesso r is important f o r sever al re asons. Firs t, the risk assessor should ensure that toxic effects thresholds calculat ed from an assessm e nt are not lower than the nutrition a l requ irem ents f o r the particu l ar p l ant or anim al species b eing evaluated. As discussed in Sections 4 and 6, the risk assess or should be aware that such elem ents exhibit classic bell-shaped or biphasic dose-response (or exposure ef fect) relationships, with adverse effects occurring at both high and low concentrati ons and an optim al mi d-range dose or exposure (see Abernathy et al., 1993; Chapm a n and W a ng, 1998). For aquatic organism s, infor m ation about nutritional requirem ents is available fo r m a ny commonly tested or cultured species. Essentiality issues also impact the bioaccumu lati on and toxicity of m e tals since org a nism s have evolved various m echanism s to m a intain hom eost asis of essen tia l m etals. Such m echanism s m a y also i mpact the bioaccumu lati on and toxicity of nonessential m e tals, particularly those that share s i milar binding an d uptake m echanism s. In these cases, accumu lation is non lin ear with respect to exposure concentration, w hereby greater uptake and reten tion of m e tals occurs at low concentrations and uptake rates decrease as expos ure m e dia concentrations increase. The impact of hom e ostatic and o ther m echanisms on bioaccu mulation has been discu ssed in Section 5.2.

5.3.2. Toxicokinetics/Toxicodynamics (Toxicity Issues) For organo m e tallic co mpounds such as organo -selen ium and m ethyl mercury, toxicity from dietary exposure has been shown to cont ribute substantially to ecological risk at environm entally-realistic concentra tions and thus should be considered by the risk assessor when characterizing the effects of these compounds. Beyond those two organo-m etal comp ounds, however, the importance of exposure to dietary m e ta ls is mu ch less clear. Toxicity to aquatic organism s from dietary exposure to m etals has b een dem onstrated where exposure is sufficiently high, although, in som e cases, these concentr ations are extrem e (e.g., 10,000 g/g Cu) (Handy, 1993). In su ch cases, it is not clear that th is path way will drive ecolog i cal risk, as th e environm ental concentrations necessary to pr odu ce thes e exp o sures m a y be so extreme that ecological risk will occu r first v i a other pathway s (e.g., direct toxi city of waterborn e m etal). In other studies, however, e ffects from dietary exposure ha ve been dem onstrated at relatively low exposure concen trations (e.g., zooplankton studies by Hook and Fisher, 2002,
5-14

### 2001a, b). This raises addition a l co ncern for m e ta ls assessm ent becau se it in creases the potential for toxicologically significant expos ures to occur in cases where risk via a waterborne pathway is low. Howe ver, other studies with the sam e or ganism s and m e tals, but som e what different test
m ethods, reached different conclu sio n s regard in g the sign ificance of dietary exposure (e.g., De Schamphelaere and Janssen, 2004; Meyer et al. [in press]). Dietary exposure of aquatic organism s to m e tals is an active area of research, and it is lik ely thatne w data and insights will result in a m ore comprehensive understanding of di etary effects. Until that tim e, risk assessors should m ake decisions regardi ng potential risks of dietary m e tal exposure on a case-by-case basis.

5.3.3. Metal Mixtures Mixtures of m e tals (including m e talloi ds and organic substances) are comm only encountered in the natural environment as a result of anthropogenic inputs and should be considered b y the risk assessor for all assessm en ts. Meta l inte ractions, ac c o rding to Ca lam ari and Alabaster (1980), occur at three levels:

• Chemical in tera ction s with oth er con s titu ents in the m edia • Inter actions with the physiological processes of the organism , and • Interactions at the si te of toxic action. The joint action of m etal mi xtures m a y be expressed in different ways, including increasing or decreasing the toxicity relative to that pr edicted for individual components. As a result, the to xicity of m etal mi xtures has important consequences for m e tals risk assessm ents. For example, toxicity has been observed for mi xtures of m e tals presen t individually at nontoxic levels (e.g., at levels correspondi ng to water quality guidelines) (Enserink et al., 1991; Spehar and Fiandt, 1986). Despite the importance of consid ering the effects and mi xtures of m etals to aquatic organism s, risk assessors w ill find thatpredicting the toxicity of m e tal mi xtures has proven to be a difficult cha llenge in aquatic toxicology. Much of the difficulty in predicting the toxic ity of m e tal mi xt ures to aquatic organism s results from differences in the bioavailability and/or m e thods used to define the bioavailable fraction am ong toxicological studies and subsequent am biguity in interpreting mi xture toxicity data. As discussed in Chapter 2, the bioavailabil ity of m e tals depends on a suite of factors that can af f

ect th eir spe ciatio n, comp lexation with lig ands, and interaction wi th biological system s (e.g., pH, DOC, inorganic anions, and cations). Ap art from bioavailability differences, the joint action of m e tal mi xtures to aquatic organism s (i .e., antagonism, additivity, synergism ) has been reported to depend on other aspects of toxicity test design, including th e degree of toxicity associated w ith the overall mi xture concentr ation (Mowat and Bundy, 2002; Fargašová, 2001; Herkovits et al., 1999; S pehar and Fiandt, 1986), the relative proportion o f constituen t 5-15

### concentrations (Norwood et al., 2003; Sharm aet al., 1999), the duration of the exposure (Marr et al., 1998), and several other factors related to experim e ntal design (Norwood et al., 2003). In a similar review, the European Centre for Ecot oxicology and T oxicology of Che micals concluded that the acu te and chron i c toxic ity to aquatic organism s of mi xtures of m e tals could not be reliably predicted or generalized although they recommended that, in the interim , assumi ng additive effects is lik ely “a balan ced approach” f or acute toxicity of m e tal mi xtures (ECTOC, 2001). Given these dif f i cultie s in evaluating mi xtures ef f

ects, ri sk as s essors comm only use two simp lifying models: concentrati on addition and effects (response) addition. These models are used to class i fy the com b ined effects of chemi cal mi xtures as being an tagonistic, additive, and synergis tic ( also ref e rredto as “ l ess than additive ,” “stric tly a dditiv e,” an d “m ore than additive,” respec tive l y )

. Both m o dels use m eta l concen tr ations in m e dia to generate concentrationresponse curves for individual m e tals, and these data are then used to gene rate specific critical concentrations for mi xture m odels. In the concen tration add i tion m odel, all m etals in a mi xture are added together to predict toxicity; differing potencies are taken into account by converting chemical concentration s to an equito xic dose, su ch as toxic units (TUs) or toxicity equivalence factors (TEF s), which converts all m e tals to one m e tal concentration. Concentration addition is often used when the cons titu ents are known or assu m ed to act through th e sam e or simi lar MOA. However, risk assessors should be aware thatapplying the concentration addition m o del to
mi xtures contain i ng m a ny m etal co nstitu ents (p articularly those well be low toxic le vels) c an result in an upward bias in predicted mi xture toxicity (Newma n et al., 2004). In the effects addition model, differing potenci es are ignored, and the effect of each
m etal’s conc entration in a mi xture is com b ined to predict mi xture toxicity. The effects addition model is use dwhen constitu ents a ct o r are assum e d to act inde pendently (i. e., dif f

erent MOAs). Thus, the risk assessor defines the nature of the m e tals’ joint action (i.e., independent or similar MOA) to decide when to apply either the concen tration addition or effects addition m o del. The risk assessor should consult inform ation on the MOA, capacities to act as analogues for other
m e tals, essentialities, and ligand binding tendenc ies to choose am ong these types of models. The assump tion of additivity has som e regul atory preceden ce for use when address i ng the toxicity of mi xtures although not necessarily for m etals. For exam ple, the concentration add ition approach is recomm ended for use by Australi a and New Zealand for evaluating whether a
mi xture of less than s i x constituents exceed s their water qu ality guidelines (ANZECC and ARMGANZ, 2000). Similarly, additivity is assu m e d by EPA when evaluating the combined acute tox i city of mu ltiple toxic ef f l uents on th e b asi s of whole-ef f l uent to xicity data ( U .S. EPA, 1991). Additivity is no t assum e d for chronic exposures due to lack of supporting data. For predicting the direct toxi city of mi xtures of cationic m e tals in sedim e nts to benthic organism s, EPA uses the ∑ SEM-AVS approach described previously (see Section 5.2.5 Toxicokinetics and Toxicodynami cs (Bioavailab i lity an d Bioaccumulation). No te, however, that this m e thod is 5-16

### limited to six cationic metals (Cd, Cu, Pb, Ni, Ag, and Zn). Furtherm ore, this m e thod is considered a “no effect guideline,” w hereby the absence of to xicity canbe predicted reliably (when ∑ SEM<AVS) but the occurrence of toxicity (when ∑ S E M>AVS) cannot be because of other factors that are not accounted f or, which reduce m e tal toxicity. Risk assessors should also c onsider the QICAR approach (d escribed in S ection 3.1.1) for addressing the toxicity of m e tal mi xt ures. Un satisf i ed with th e quali tative conclusions of Newm an and McCloskey (1996), Ownby and Newm an(2003) fit binary m e tal mi xture data derived from the Microtox assay to develop a m odel of joint i ndependent action (Finney, 1947). They predic ted that the joint ac tion o f com b ined m etals will increas ingly deviate f r om independent action as their ligand-binding chemistries becom e m ore and more similar. Although Micro t ox is consider ed to be a usef ul tool f o r org anic con t aminants, its se nsitiv ity f o r evaluating
m etal toxicity has been called into qu estion (W illem son et al., 1995). It is po ssib l e that recep to r binding m odels (e .g., FIAM) m a y be expandedin the future to include mi xtures. In theory, if two m e tals comp et e for binding to the sam e site of toxic action, it should be possible to m o del the to tal m e tal bound to that site and, he nce, to predict m e tal toxicity using a m echanistic receptor bindin g approach inan effects additi on m odel. Alternatively, if two
m etals do not com p ete for the sam e binding site, then these m odels m ay provide m o re reliable estim ates of individual metal bioavailability, wh ich then can b e com b ined in m o re accurate effects addition m odels. However, at presen t, th ese possib i lities rem ain theoretica l. Furtherm ore, this approach, while improving the ab ility to ass ess the ef f

ec t s of m etal mi xtures, does not include temporal aspects (i.e., “tim e-to-response” versus concentration). From the preceding d i scussion, it sh ould be clear that the accurate pred iction of the joint toxicity of m e tal mi xtures to aquatic organi sm s rem ains a significant ch alleng e for the risk
assesso r. For site-specific assessm ents, risk asse s sors are encouraged to assess mi xture toxicity using in s itu m easurem e nts (i.e., bioassays using site water or sedim e nts). This approach is the foundation of the W E R procedures used by EP A for m aking site-specific bioavailability adjustm e nts to m e tals criteria. For site-specifi c assessm ents involving sedim e nts, risk assessors should consider using the ∑ SEM-AVS approach as a no-effect threshold. For national-level
assessm ents, there is som e preced ence for assumi ng additive toxicity of mixture constituents, particularly when considering acut e effects. However, the risk assessor should caref ully consider the limitatio ns to assumi ng strict add itiv ity (i.e ., potential for overpred i ction or underp rediction of toxicity) and highlight these uncertain ties in th e Risk Char acte r ization phase of the aquatic risk assessm ent.

5.3.4. Critical Body Residues The bioavailability of m e tals from mu ltip le expos ure routes (water co lumn, food, sedim ents) should be co nsidered in aquatic risk assessm ents to account fo r relative co ntribu tions to overall toxicity. In concept, expressing toxicity on the basi s of tissue residues is an attractive 5-17

### approach to accom p lishing this be cause it integ r ates ch emical uptak e from different routes of exposure, accounts for differences in bioavailability from ex posure m edia, 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 relations hips) has gained significant attention in the aquatic ecotoxi cology literature, particularly for organic chemicals (e.g., Landrum et al., 2005, 2004; Escher and Herm en s, 2002; McCarty and Mackay, 1993; Cook et al., 1993, 1989; McCarty, 1986; Veith et al., 1983; Könem a nn, 1981). For m a ny nonionic organic chemicals, available data indicate that whole-body burdens of chemical (normalized to lipid content) can serve as useful m e trics of toxi cological dose, and these relationships appear to be independent of whether exposure was via water or diet. Am a jor strength of the CBR approach for organic chemicals is thatit ef f

ectively integ r ate s di fferent exposure pathways into a single expression of dose a nd toxicological potency. For m e tals (aside from organo-selenium and m e thyl m e rcury), the situation is far m o re complex and the CBR approach does not appear to be a robust indicator of toxic dose. One reason why the CBR app r oach curren tly appears more limited for m eta ls relates to differences in their m echanism s of upt ake, dis tribu tion, and d i s position in aquatic o r gan i sm s. Specifically, the distribution of nonionic organic chemicals in organism s is largely influenced by passive partition i ng. In contrast, the uptak e, distribution, and disposition of m e tals are typ i cally governed by highly-specific biochemi cal proce sses that alter the m e tal form and involve facilitated or activ e trans port. For ex ample, so m e organism s take up m etal and sequester it into “storag e” co mpartm ents in chemical for m s that have little toxicological potency, whereas other organism s actively excrete excess metals. As a basis for improving residue-response relationships, som e studies have sugg ested that th e m etal concentr ation in the cellular c y tosol ( as opposed to that bound to cell walls or sequest ered in nonbioavailable m e tal granules) m a y provide a better expression of internal m e tal dos e associated w ith tox i c effects (W allace and Luom a, 2003; W allace et al., 2003 ; Wallace andLopez, 1996). Other researchers have suggested that CBR relationships are confounded because the factor that d etermines th e effects is n o t whole-body concentration per se, but the ra te of m e tal uptake in relation to m etabolic capac ity f o r detox i fication andstorage; th erefore, the effects are governed by factors that influe nce the rate of uptake. When uptake rate is elevated, the concentratio n of m e tabolically active m etal at the s ite(s) of action increases (e.g., the s p illov e r hypothesis) and effects ensue (Rainb ow, 2002). B ecause different sp ecies of aquatic organism s invoke different “accumulation strategies” (i .e., involving co mbinations of regulatio n, detoxification, and storage), cons iderable difficulties arise among species when interpreting the toxicological significance of m e tal whole-body residues. Therefore, risk assessors should ensure thata toxicologically valid residue-response relationship supports the CBR thres hold before using tissue residues as indicators of toxicity. 5-18

### Although many toxico lo gical studies report m eas urem ents of m etal residu es in mu ltiple tissu es along with adverse effects, these tissue residue values m a y not be appropriate for use as a CBR threshold because m e tal concentrations in som e tissues m ay have little or no relationship with toxicity. Furtherm ore, risk assesso rs are cau tioned agains t ex trapolating C BRs across differing exposure ro utes (food v s. water), du rations, tissu es , or species, because th e potency of m etal residues often differs de pending on these factors.

5.3. RISK CHARACTERI Z ATION As described in Section 2.4, Risk Characte rization is the final phase of the risk
assessm ent and is the culmination of the Plan ning, Problem Formulation, and Analysis of predicted or observed adverse effects. Risk Ch aracterization produces a d e tailed description of the risk estim a te(s), evaluates and summ arizes the lines of evidence that support or refute the risk estim ate(s ) , describ es th e uncertain ties, assump ti ons and qualifiers in the risk assessm ent, and reports the conclusions of the assessm e nt to risk m a nagers (U.S. EPA, 2000c, 1998a). While there are no m e tal-specific m e thods in the Risk Characterization, there are aspects that ar e importan tto m etals ri sk ass e ssm ents. For example, considering multiple lin e s of evidence su ch as resu lts from in situ toxicity testing or biological assessm e nts can be valuable for supporting the conclusions of a risk assessm e nt. Care should be taken, however, to evaluate and presen t the limitation s associated with each line of eviden ce, as dis c repancies m a y not always ind i cate underlying differences; rather, th ey m a y reflect inherent limitations of each of the m ethods. For exam ple, biolog ical assessm ent m ethods m a y not be sufficien tly s ens itiv e to detect the level of effects or exposures that are of concern in the risk assessm ent. Docum e nting
assumptions and uncertainties (e.g., use of bac kground m e tal concentrations rather than added
m etal or spe cif ic m etal s p ecies ) beco m es increasi ngly important the closer hazard thresholds are to background concentrations. Risk assessors also should docum e nt all assumptions and uncertainties in the m e thods used, su ch as how metals bioavailability was addressed. Because data m a y not be adequate to conduct a quantita tive uncertainty analysis , risk assessors should describe the sensitiv ity of the risk assessm ent results to key assump tions and the direction of bias introduced by these assumptions (i.e., under- or ov erestim a tion of risks). This is particularly important for national or regional assessm ents, where results often are intentionally based on organism s and condition s that enhan ce exposure, bioavailability, and toxicological sensitiv ity. For essential m e tals, risk assesso rs sh ould descri be the relationship of the risk thresho l d to nutritionally required levels for the organism s of concern. Ris k assessm ent resu lts that f all below nutritionally required levels are an indication that som e m e thods or assumptions require additional refinem e nt. Risk asse ssors should carefully document th e form (s) of m etals used in the exposure and effects assessm ent, as they freq uently d i ffer due to data limitations. Additional issues and questions that should be addressed in the risk asse ssm e nt are listed in Section 2.4.

5-19

### 6. TERRE STRIAL E C OLOGICAL RI SK ASSESSMENT FOR M ETALS

This section of the Fram e w ork provides an overv iew of how t he principles for m e tals risk
assessm ent apply to eco l ogical risk assessm ents fo r terrestrial environm ents

. Receptors typically considered in these as ses sm ents inclu d e soil inve rt ebrates, p l an ts, and wildlif e spec ies

. Som e assessm ents also examine effects on mi crobiota and soil processes. This section of the Fram ework builds on the inform ation presented in Chap ter 2 that lays out issues to be considered during Problem For mulation and that describes me tal chemistry assoc i ate dwith soil s y stem s. That information is not repeated here and th e reader should refer to Chapter 2 for this inform ation.

6.1. METALS PRINCI PLES Metals have specific environm ental and biotic attributes that should be considered in all risk as sessm ents. These principles fo r m etals risk assessm ent (see Chapters 1 and 2) apply in various ways to ecological risk assessm e nts depending on the scale of the assessm ent (site specific, regional, or national). Th is section describes applications of the princ i ples to terr estria l ecological assessm ents within th e standard risk assessm ent fram ework. Specifically, they fall into th e risk assessm ent paradigm as f o llows:
Background levels Exposure Assessm ent Mixtures Exposure and Effects As sessm ent Essentiality Effects Assessm ent Form s of m e tals Exposur e and Effects As sessm ent Toxicokinetics/toxicodynamics Exposur e Assessm ent (bioavailability) and Effects Assessm ent (ADME and toxicity )
6.2. CHARACTERI Z ATION OF EXPOSURE Metal exposure assessment include s characterization of the ex posure routes and pathw a ys specific to m etals, the p h ase associations and ch emical for m s of the m e t a ls, and the expression of exposure an d target doses in a m a nner consisten t with definin g hazard thre sholds for particular organism s.
6.2.1. Natural Occurrence of Metals At a nation a l leve l, m etal concen trat ions vary naturally in soils across the U.S. These variations pose challenges for conducting national assessm e n ts of risk to terrestrial ecological recep tors. T h e assesso r m ay decide to use a sing le toxicity level reg ardless of background concentrations for a screening type assessm ent (see text box on eco logical soil screening levels or Eco SSLs) (U.S. EPA, 2003c) or m a y prefer to divide the country into regions of similar m e tal
6-1

### background levels (metalloregions). Exposure assessments should consider metal levels inclusive of background. At the regional and local (site) scales, risk assessors should accoun t 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 survey 4 

### is over 20 year old, risk assessors should consider the feasibility of generating site-speci fic concentrations for local risk assessments. s. s

More appropriately, risk asse ssors should avoid single-result assessm ents for the entire country. Rather, such assessm ents should be su bdivided into m e tal-related ecoregions known as “m etalloreg i ons” (McLaughlin and S m olders, 20 01) so thatp r otec tion lev els, mitigation goals, and ranking results will be appropriate for the su ite of species natura lly p r esent within each type of controlling environm ent. This is direc tly analogous to the us e of ecoregions when establishing water quality criteria (Griffith et al., 1999). The use of m e talloregi ons provides the ability to account for the broad regional param e ters affecting m e tal ava ilability in soils and waters as well as for the differences in organism response to added m e tal. 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 system s (Bailey et al., 1994; Ba iley, 1983). These are based on climactic and vegetation factors and form the basis of metallore gions. EPA is still working to complete ecoregion maps at much fi ner scales for each state (see EPA Web site at http://www.epa.gov/wed/pages/eco regions/ecoregions.htm )

. To complete the metalloregion concept, soil properties that a ffect bioavailability (e.g., pH, cati on exchange capacity [CEC], and organic matter [OM]) should be overlaid on th e ecoregions, along withsoil 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 fa irly current and availabl e, 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., Mc Laughlin 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 Schac k l e tt e , H T ; B o ern g en , J G

. (1 9 8 4 )

. El e m ent conc entr ati ons i n s o i l s an dsu rfi ci al m a t e ri al s of th e c ont e r mi nous Uni t e d St at es. U.S . Geol ogi ca l Su rvey Pr ofe ssi onal Pape r 1 2 7 0

. 1 0 5 pp . EPA’s Ecological Soil Sc reening Lev e ls EPA’s Ecological Soil Sc reening Lev e ls (Eco SSL s) f o r metal s are national -level c o nc entrations of metals in s o ils that are prote c tive of wildlife, plant s, and soil orga nisms. T hese value s are lo we r tha n naturally o c curri ng level s in some p art s of the cou n try. Exceede nces of such level s doe s not mea n that a risk e x ists but doe s mean that a more regio nal or site-sp e cif i c asse ssm ent may be need ed.



### 6-2

### 6.2.2. Forms of Metals The physical and chemical form s of m etals influence exposure and subsequent effects and can be influenced b y physical/chemi cal c onditions in th e environm ent. National level
assessm ents involve a broad range of environm en tal conditions and so the risk assessor should account for different m e tal species in different locations and soil types. As assessm ents trans i tion f r om national, to reg i onal, to loca l, the assesso r sho u ld inco rpor ate s ite -spec i f i c soil param e ters that influence m e tal speciation (e.g., pH, CEC, cl ay content). National values (e.g., geom etric mean values) of these param e ters s hould be used, with the sam e recomm endation as discussed in Section 5.1.3 on Na tural Occurrence of Metals.

6.2.3. Exposure Routes The m ajor m etal exposure route th at the risk as sessor shou ld consider f o r wildlif e is ingestion, with a mi nor (and often unknown) inhalation comp onent. For plants, root uptake is the m ost important with leaf exposures secondar y, with the exception of Hg where the m a jority is accumulated via foliar uptake; Cd and som e ti me s Pb also may be accumu lated throu gh foliage but am ounts relative to s o il exposu r e will vary de pending upo n soil cond itions (e.g., p H ). Plants
m a y also lose m e tals through foliar leachi ng du ring precip itation events although to a significantly lesser extent than for other mi cronutri ents such as potassium. Soil invertebrates are
assum e d to be exposed through direct contact. Pa th way s describe transport of the co ntaminant in the environm ent and include uptake and bioc oncentration (e.g., dietary ingestion of a soil contaminant that has been taken up by plants). Princip l es of m etal transp ort and f ate in soils ar e applic able to assessm ents of risk to all ter r estria l organism s and will be d i scussed f i rs t. However, because of sig n ificant d i fference s in exposure rou t es and path ways between invertebrates, plants, an d wildlife, it is m o re convenient to discuss exposure assessment m e thods by recep tor group.

6.2.4. Soil Transport and Fate Models Path w ay of Exposur e fo r 

### Risk assesso rs routinely use transport and fate Terres t rial O rganisms 

### models (i.e., a computational m odel) to describe and Pathw ays o f expos ure fo r 

### quantify exposure pathw a ys. Models are also useful in terre strial or ganisms to metals inclu de move ment from soi l s throu gh 

### situations w here risk assessors are trying to estim a te the food we b, and to a lesser extent, air dep ositio n either into so ils or 

### exposure levels that are expected to result from the dire ctly onto terrestri al re ce ptors (e.g., 

### implem entation of som e permitting a ction o rrem ediation plants).


### m easures at local, regional, ornational scales. N u m e rous models are availab l e for use; m o st are based on the sam e fundam ental principles: m e tals are ubiquitous in the environm ent and within ea ch m edia comp artm ent they are present in association with water ( f r eely disso lved m etal oras organic and inorganic m e tal comp lexes),
6-3

### particles (sorbed, precipitated, or incorporated with in a mi neral phase), and air. The risk assesso r can find a more detailed discussion of these pr ocesses in Section 3.2 on Fate and Transport. Currently, th ere is no sin g le m odel availab l e th at encom p asses all the des i rable m etal-specific features for terres trial sys t em s. Discussions of the available terre strial transport and fate m odels, as well as a num b er of chemical equilib rium m o dels, m a y be found in Allen (2002).

6.2.5. Toxicokinetics/Toxicodynamics Target organ exposure levels and subse quent effects depend on how environm ental conditions affect speciation of a m e t a l (e.g., whether an organism actively takes up or excludes
m e tals in soils and how an organism processes me tals interna lly). Se e Se ction 3 f o r d e tails on environm ental chemistry and issues relating to b i oaccumu lation. Risk assessors shou ld specifically address b i oavailab i lity and bioaccu mulation for each m e tal of concern in each environm ent (either a local site for site-specific as sessm ents or som e larger estim ate for regiona l and national level assessm e nts).

6.2.5.1. Bioavailability Risk assessors should adjust bul k soil m e tal concentrations by appropriate bioavailability f

actors to ac hieve comp arable, actua l uptake of me tals by soil organism s. This will standardiz e exposure values across soil typ e s andallow for more accu rate comp arisons with labo ratory toxicity data. Cation exchange capacity (CEC) recently h a s b eenshown to be an important factor m odifying zinc bioavailab ility in soils, and presum ably it will be important for other cationic m e tals as well. However, CEC is str ongly dependent on the type and am ountof organic
m aterial (O M) and oxyhydroxides present in the soil, and is strongly pHdependent. Surface charge on O M and oxyhydroxides incr eases with pH, thereby increa sing their sorptive capacity for m e tals (thus decreasing m e ta l bio availability ). Conversely , positiv e su rface charg e s increase as the pH drops, which increas es sorption of an io ns (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 indep e ndent of pH. Therefore, norm alizatio n of toxicity data to CEC can be done only within specific soil type s and pH ranges, which frequently are not specified either in laboratory bi oassays or m a ny field studies. Fu rtherm ore, it is important for the risk assessor to note that m ost published valu es of CEC ar e m easured at pH 7. In general, risk as sesso rs can assum e that cation i c m etals are m ore bioavailable at lower soil pH (<6) and less bioavailable at higher soil pH (>8). Th e opposite assumption holds for anionic metals. Soil chemical m odels ar e being deve loped to p r e d ict how aging will m odif y bulk soil concentratio ns when soils are am ended with so luble salts. A ging reduc es the bioav ailable fraction of m e tals over tim e. Preliminary studies suggest that considerat ion of aging m a y result in estim a tes of the bioavailable fraction as low as 0.1 × bulk soil concentrations (McL aughlin et

6-4

### al., 2002). Until th e data becom e availab l e for meta ls of concern, tox i city values deriv e d from soluble-salt am ended soils (which have not simu lated aging) cannot be reliably corrected to approxim a te aged m e tals in field situations a nd the risk assessor shoul d acknowledge this as a significant uncertainty during the risk characterization. Ideally, exposure should be expressed on th e basis of pore water concentration, to account for all factors influencing bioavailabil ty; however, there are currently significant limitations to colle cting and inte rpre ting m etal-re lated d ata f r o m soil pore waters and such inform ation generally is not available (even at s ite-specific assessm ents and never for regional or nationa l ass essm ents). The risk as s essor cou l d estim ate m etal concen tration in so il p o re wate r using EqP theory (as with sedim e nt pore water analyses; see S ection 3.1.5 ). The risk assessor can use published soil binding coefficients (Kd 

### s) to estim ate partition i ng b etween soil particle s and pore water although these values also are inherently uncerta in (published value depends on derivation m e thod, soil type, etc.). Furtherm ore, toxicity threshold values generally are provided as bulk soil concentrations so the risk assessor would not be able to co mpare pore water exposure with any effects es tim ates.

6.2.5.2. Bioaccumulation For terrestrial ecosystem s, the concept of bioaccumulation is intended to capture the potential for two ecologically important outcom e s: (1) direct toxicity to plants and wildlife and (2) secondary toxicity to anim al s feeding on contaminated plants and an imals. This ap proach stresses the potential for trophic transfer of m e tals through the food web, so to tal exposure can be calculated, including dietary intake as well as intake from cont aminated environm ental m edia (soil and water). Fo r vegetation or s o il inv erteb rates, the bio accumu latio n factor (BAF; or biotasoil accumulation factor, BSAF) is defined as fiel d m easurements of m e tal concentration in plant tissues divided by m e tal concentration in soil (or soil so lution ); th e BCF is defined as the sam e m easurem ent carried ou t in the laboratory (Sm olders et al., 2003). Risk assesso rs should be aware that d ata app li cab ility is directly related to which tissu e is
samp led and how it is processed. B A Fs for plants include metals aerially de posited o n leaves as well as thos e in soil particle s adheri n g to roots. S u ch m e tals will not b e p art of BCFs, which frequently are determined in hydroponic culture. Similar differences betw een BCFs and BAFs apply f o r earthworm s exposed in soils versus lab o ratory stud ies using th e f ilter p aper s ubstrat e protocols. Furtherm ore, BCFs within earthwo rm s m a y not include addi tional feeding of the anim als during the study. Field studies are refl ective of chronic exposures, whereas BCFs m a y be calculated from shorter tim e fra m es. Ideall y, risk assessors should select BCFs reported at equilibrium (i.e., after sufficient exposure tim e to m a xi mi ze the BCF). Wh ole-body B A Fs generally are not calculated for birds and m a mmal s, except for sm all m a mmals such as rodents (Sample et al., 1998b). Risk assessors should understand the conditions under which m e tal concentratio ns were m easured and critically examine data to determine whether th ey are reporte d 6-5

### as wet or dry weight (the ratio of tissue to soil concentrations mu st be done on the sam e wet/dry weight basis for both). For soil inve rtebr ate s an d m o st plants, m e tal BAFs are typically less than 1 and usually are bas ed onthe to tal m etal in soil an d tissue th at do not accou nt for bioavailab ility differences. The risk assessor mi ght consider using a ratio of tota l m etal in the organ i sm to som e m easure of the bioavailable fraction of m e tal in the soil (e.g., free ion concentration or weak salt extractable) for expressing a BAF to allow comp arison am ong di fferent soils, although, in general, data are lacking for using this m e thod. Furtherm ore, the risk ass essor is remi nded that b i oaccumu lation of m etals is not a s i mple linear relationship. Uptake is nonlinear, increasin g at a decreasing rate as m edium concentration increases. Models for predicting m etal b i o accu mulation by soil inv erteb rates are p r im arily statistical in nature, describi ng relationships betw een m e tal body burdens in oligochaetes and collem bola, soil m etal co ncentr ation s , and soil ph ysica l/chemi cal char ac teri stics. Samp le et al. (1998a) and Peijnenburg et al. (1999b) have each developed un ivariate uptake m odels for earthworm s that are bas e d on em pirical data (m etal concentrations in worm s vs. the natural log of a m ount of m e tal in soils) that risk assessors can use as a first approximation for bioaccumu lation in so il inverteb rates; however, th ese m odels are not sp ecific to so il ty pe and, therefore, do not account for bioavailability factors. Furthermore, they do not adequately predict Cr or Ni uptake. An alternative approach thatthe risk assesso r could consider is the use of multivariate statistica l models to loo k f

o r patte rn s of uptake of mu ltiple m etals to pr edict the potential bioconcentration of one m e tal of particular interest (S cott-Fordsmand and Odegard, 2002) or BAF as a function of soil characteri stics (Saxe et al., 2001; Peijnenburg et al., 1999a, b). Path analysis has been sug g ested as an alternative for mu ltiple regress i on in describing these relationships. It partitions simpl e corr elation s into direc tand indirect effects, providing a num erical value for each direct and indirect effect and ind i cates th e re lative streng th of that correlation or causal influenc e (Bradham , 2002; Basta et al., 1993). The absolute level of metal accumu lation is not as im portant as the rate o f uptake (Hook and Fisher, 2002; Hook, 2001; Roesijadi, 1992). Adve rse effects are avoided as long as the rate of m e tal uptake does not exceed th e rate at which th e organ i s m is able to bind the m etal, thereby preventing u n acceptable increases in cytosolic lev e ls of bioreactive form s of the m e tal. If the rate of uptake is too great, the comp lexation capacity of the binding ligand (e.g., m e tallothionein) could be exceeded; cy tos o lic m etal levels th enbecom e unacceptably h i gh , and advers e effects
can ensue. Because m e asures of uptake rates are notavailable, static con centration s are used; the risk assessor should acknowledge this uncertainty during the Ri sk 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 oneach other, decaying p l ant m aterial, and bacteriaor fungi. However, 6-6

### the risk assessor should estim ate exposure as a fu nction of soil concentration, rather than as a detailed analysis of m ove m e nt of metals thr ough the food web, to generate data that will be
comparable to effects concentrations. This is a reasonable approxim a tion for soft-bodied invertebrates (e.g., earthwor m s) whose m e tal expos ure is prim arily through soil pore water (from both derm al absorption and soil ingestion) (A llen, 2002). There is m ore uncertainty in
corre lating s o il m etal co ncentr ation s with ef f

ects in hard-bod ied inve rteb r ates be cause they ar e prim arily exposed through ingestion of food and incidental amounts of soil (Sample and Arenal, 2001). Regardless, risk assessors should estim a te soil invertebrate exposure on the basis of total
m etal concentration in b u lk soils (ad j usted f o r r elative b i oavailab ility, where poss i ble ) colle cted in the top 0-12 cm of soil (U.S. EPA, 2003c, 1989b). In detailed, site-specific assessm e nts, the organic m a tter on top of the soil (the “duff”) m a y be analyzed separately to provide further detail on exposure to detritivores (such as Coll embola ) and deeper-soil-dwelling organism s (e.g., various species of earthwor m s).

6.2.7. Plant Exposure Plants access m e tals thro ugh the pore wa ter altho ugh my corrhyzae, pro t o n s, and phytosiderophores released by the root can si gnificantly influence the mi croenvironment and change uptake rates of m e tals (George et al., 1994; Sharm aet al., 1994 ; Laurie and Manthey, 1994; Arnold and Kapustka, 1993). Furtherm or e, plants have both active and passive
m echanis m s for taking up or excluding m e tals, de pending on internal con centrations and whether or not the metal is an essential mi cronutrient, or whether it is mi staken for an essential
mi cronutrient. Plants can be exposed to m e tals via aerial deposition o nto leaf surfaces, trapp i ng m e tals in hairs or rough cuticu l ar su rfaces. This mi ght provide an exposure route for herbivores; it m a y also provide an exposure route for plants, as there are ion channels through the cuticle that are ab le to transport ionic m etals from the leaf s u rface to oth e r loc ation s in the p l ant, depending on the inherent m obility of the m e tal in the xy lem and phloem (Mars c hner, 1995). The risk assessor should consider the default approach to estim a ting exposure of plants to
m etal as m e asuring m etal concen trations in b u lk soil (top 0-12 cm ). However, as with soil invert ebrate s, this ove re stim ates exp o sure because it does not account for differential bioavailability and aging. The risk assessor ge nerally can categorize m e tal bioavailability and uptake based on soil pHand organic m a tter (see Sec tion 3.1.6.5). It is very clear that strongly acidic soils increas e plan t 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 significan t extent at any pH (Chaney and Ryan, 1993).

6-7

### Table 6-1. Qualitative bioavailability of metal cations in n a tural soils to plants and soil inverteb rates

Soil type Soil pH Lo w 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 inverteb rates

Soil type Soil pH Lo w 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 r elationship s between soil chemistry and bioavailab ility ar e ap propriate f o r national-scale appli cation. However, for site-specific or m etals-spe cif i c applications, the ri sk assessor should use quantitative m e thods. Parker and Pedler (1997) and Lund (1990) have suggested that only uncomplexed, fr ee ionic species of cations can be taken up by roots, and this has been described using a Free Ion Activity Mode l (FIAM) similar to th e Biotic L i g and Model (BLM) used in aquatic system s. However, si gnificant exceptions to the f r ee-ion m o del have been identified; so until this theo ry is tested m ore thoroughly, th e risk assessor should continue to estim a te exposures using bulk soil values withqualitative estim ates of bioavailability based on soil type (pH and OM). Again, the risk assesso r should acknowledge these uncertainties during the Risk Ch aracterizatio n.
6.2.8. Wild life Exposu re The relative im portance of exposure pathways and routes varies by anim al species and by
m e tal, although, in general, wild life exposure is prim arily through diet and incidental ingestion
6-8

### of soils or sedim e nts. There are certain chemicals and exposure situations for which i nhalation or derm al pathways are important, but in m o st situations the risk assessor can consider the mto be insignif i cant co ntribu tors to tota l m etal loads (U.S. EPA, 2003c ). W ildlife food chain expo sures for m e tals are controlled by bioavailabili ty, bioaccessibility, and bioaccumulation. Bioaccessibility of m e tals to anim als and plants that live on or in th e soils can be influenced by soil param e ters, such as pH, CE C, and organic carbon. These soil param e ters tend to be le ss important f o r soils th atare incid en t ally inges t e dby wildlif e species. The relative im portance of exposure pathways (soil vs. food chain) is dictated by the fraction of m e tal-contaminated soil in the d i etand the am ount of accumu lation of metal in food item s . In the absence of site-specific info rm ation, the risk assessor can use the following genera lizat ions to determine the relative im portan ce of incidental soil ingestion versus dietary
m e tals: Exposure Pathway for Terrestrial Wildlife Food and the incidental ingestion of soil are the two most important exposure pathways for terrestrial wildlife.



### 1. Inciden t al s o il ing estionis a proportionally m ore important pathway for herbivores than for carnivores or invertivores.
2. Uptake into soil invertebrates (e.g., earthw orm s) is a proportionally m o re im portant pathway for anim als that feed on thes e organism s. (Note: This assessm ent reflects work done with earthworm s and may not apply to hard-bodied soil invertebrates such as Colembolla .)
3. If bioaccumulation is lo w (<<1), importanc e of soil ing e stio n versus diet for m e tal exposure increases.
4. When bioaccumulation is greater (~1 or higher), the food pathway should domi nate. 5. The closer the association an anim al ha s to the ground, the great er the importance of soil ingestion. This associ ation m a y be due to ground foraging, burrowing habits, etc.
6. The looser the association with th e 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 contributionof food and soil before accounting fo r bioavailab i lity. The ass essor shou ldassum e that incidental ingestion of soil becom es proportionally m o re im portant for exposure to wildlife when (1) the bioaccumu lation facto r (BAF) from so il to food (e.g., to p l an ts or so il inv erteb rates) is les s than 1 and (2) th e f r action of s o il in th e diet is greater than 1%. However, the risk
assesso r sho u ld use th ese generaliz ations with caution for site-specific assessm ents. As the risk assessor acquires m ore site-specific infor m ati on, the relative importance of pathways m a y 6-9

### change. For example, site-specific data m a y s how that the accumulation o f a chemical into plants or soil invertebrates is much lower thanindicated by the default assumptions. In such cases, incidental ingestion of soil would becom e proportionally m o re i mportant. The bioavailability of m e tals in inci dentally ingested soil is also v ariab le. Therefore, when the exposure is being driven by incidental soil ingestion, the risk asse ssor should consider refinem e nts of exposure estim ates through a be tter understanding of bioavailability, although very little in f o rm ation is availabl e on this f o r m o st wildlif e spe cies.

100

80

60

40

20

0

Increasing Percent Soil in Diet BAF = 0.01 BAF = 0.1 BAF = 1 BAF - 10Percent contribution of soilingestion to total dose



### Figure 6-1. Generaliz e d representation of percent contribution of in cidental soil inge stio n to oral do se for w ildlife at different soil inges t ion rate s a n d bioaccumulation factors and a bioavailability o f 100 percent.

The risk assessor should be cautious about ex trapolating bioavailability adjustm e nts for wildlif e f r o mm odels developed f o r estim ation o f bioavailab i lity of m etals in soils f o r

inciden t al hum a n exposures. There are significant variatio ns in digestive physiology and anatomy across
m a mmalian and avian sp ecies that alter the d e gree of assimilation and uptake of m e tals (MenzieCura and TN&A, 2000). For example, m e tals presen t in soils m a y be m o re or less bioavailable within th e g u t of an herb ivore that relies on ferm entation as co mpared to the simp ler gut of a carnivore that is designed to break down prot ein s. These gut system s differ in ch emistry (including pH) and residence tim e . Food chain modeling can be used to estim ate the exposure of wildlif e to m e tals based on the ingestion of soil, food, and water. The ri sk assessor should use the sam e dietary uptake

6-10

### model for metals as is used for organic s ubstances, e.g., Eco SSLs; Ecological Committee on FIFRA Risk Assessm ent Methods (E COFRAM) (S ample et al., 1997; U.S. EPA, 1997d). For national or regional risk assessm ents, the assesso r m a y use trophic transf er rates to m o del food concentrations but only on the basi s of soil m easurem ents (rather th an using direct m e asures of concentration of m e tals in food item s ). As with aquatic organism s, trophic transfer values for
m etals in terr estria l sys t e m s are an inverse f unc tion of soil co ncentration s . 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 conc entrations and use which ever value(s) that are consistent withthe degree of conservatism or am ount of r ealism appropriate for the assessment. Samp le et al. (1998a) developed uptake m o dels to predict concentrations in earthworm s from soil concentratio ns and Efroymson et al. (2001) provides similar infor m ation for plants (see Section 6.5.2.3 Bioaccumulation for a m o re detailed discussion). W ith the exception of a few hyperaccumulator species, the risk assessor can reasonably
assum e that most plant species do not bioconcentrat e m e tals (i.e., BAFs <1). Pb, As, Cr, and Co are not taken up by plants in m easurable quantities, and the sm all a m ount that is taken up is mostly confined to root tissues (Chaney etal., 2000; McGrath, 1995; Chaney and Ryan, 1994; Xu and Thornton, 1985). In contrast, m a ny plants are quite sen sitiv e to som e m etals (M n, Zn, Cu, for example); the risk assesso r should be aware that plants fr equently die before achieving high m e tal concentration levels that pose a thr eat to anim als via food ch ain transfer (with the exception of the hyperaccumulator species, as noted above).

6.3. CHARACTERI Z ATION OF EFFECTS When assessing m e tal toxicity to terres trial organism s, the risk assessor should understand both the natural m echanism s of toleran ce for (or, in the case of mi cronutrients, the use of) m e tals and the toxicological responses th at occur when exposure exceeds the capacity of the organism to regulate its body burdens. The ri sk assessor should also consider interactions between m etals in eith er their uptake or toxicity (such as Cd/Ca/Zn, Hg/Se, Cu/Mo). Risk
assessm ents for m etals are further co mplicated b y the need to express th e dose-respo n se (or concentration-response) functions in bioavailable units that are functionally equivalent to
m easures of exposure. T h is section provides t ools and approaches risk assessors can use when addressing issues of essentiality, m e tal mi xtures, and appropriate use of toxici ty tests; issues of how acclim ation or adaptatio n to continu e d exposures
m a y a ffect toxicity have b een 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

6-11

### Essentia lity , or the requirem e nt for norm a l organism m etabolic function, of som e m et als is one of the prim ary factors that differentiat es risk assessm ent for m e tals and m e tal compounds from that of synthetic organic chemicals (J anssen and Muyssen, 2001). Som e trace elem ents, such as Co, Cu, Fe, Mn, Se, Mo, and Zn, are nece ssary for the norm a l developm ent of plants and anim als. Other m e tals, such as As, Cd, Pb, and Hg, have no known functions in plants and anim als (Mertz, 1981). Table 6-3 classifies the m e tals addresse d in this Fram ework by their known essentiality to organism s. The risk assessor should be sure that eff ects thresholds such as Toxicity Reference Values (TRVs) are not lower than th e nutrition al requir em ents f o r the particularplan t or anim al species being evaluated. If TRVs are set too low (i.e., in the range where deficiency can occur), the d eterminationof risk will be erroneous and deficien cy effects will be
mi staken for toxic respo n ses. For wildlife, th e risk assesso r can consult the literature on dietary requir em ents of essentia l elem ents f o r liv estock (McDowell, 2003; NAS/NRC, 1994a , 1980). Marschner (1995) summarizes the mi nimu m concentrations required for plant growth. Because of differences in test conditions
am ong published studies, it m a y be difficult for the ri sk asse ssor to d i rec tly comp are toxic ity thresho l d va lues withrecomm ended dietary requirem ents of essential elem ents. Extrapolation of data among species (e.g., from livestock to wildlife species) m a y also add uncertainty to the effects assessm ent. Furtherm ore, addition of safety factors when deriving protec tive v alues of ten r esults in con centr ation s significantly below required intake. The risk
assesso r sho u ld addres s these and similar uncertainties in to xicity thr esh old derivations as part of the Risk Ch aracterizatio n process. Detailed site -specific ass essm ents, where m o re accurate estim ates of ef f

ects thres holds are ex pected, m a y requir e the ri sk asse ssor to reques tad dition al bioassays to characterize the biphasic dose-res ponse curve and determine both required and excessive threshold levels.

6-12 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 additionalbioassays to characterize the biphasic doseresponse curve and determination of both required and excessive threshold levels.

### Table 6-3. Metals classified by their know n e ssentiality

Metal Essential (known requirement for

health and f unction) Beneficial (but not known to be essential) Nonessential (and not known to be beneficial ) Plants Animals Plants Animals Alu minum (Al) x Antim ony (S b) x Arsenic (A s) X Bariu m (Ba) x Bery llium (Be) x Cad miu m (Cd) x Chromiu m ( C r) x Cobalt (Co) x x Copper (Cu) X x Lead (Pb) x Manganese ( M n) X x Mercury (Hg) x Mol y bden u m (Mo) X x Nickel (Ni) X x Selenium (Se) x x Silver (Ag) X Strontium (Sr) X Thallium (Tl) X Vanadiu m (V) X Zinc (Zn) X x Source: Adapted from a tab le presented in SRWG (2 002) and i n cor porating data from NAS/NRC (1980) and Barak (1999). Fairbr other and Kapustka (1997) di scussed the r oots of essentiality of natur a lly occurring elements.



### 6.3.2. Toxicity Tests For assessm ents conducted for regional or national assessm e n ts, criteria developm ent, or ranking purposes, risk assessors should acknowledge that results will be based on organism s and soil typ es th at resu lt in g r eatest b i oav ailab i lity an d sensitivity. The risk assessor should take great care to ensure th at the organ i s m -environment com b inations th at ar e assess ed are, in fact,
6-13

### com p atible with rea l -wo r ld cond itio ns. Thus, f o r site-spe cif i c assessm ents, spec ies te sted and water (or sedim e nt) used in the test system shoul d be similar to conditions at the site. In the absence of such inform a tion, risk assessors co uld use data from sta ndard test species and conditions, but uncertainty factors m a y be warr anted to adjust the fi nal toxicity value.

6.3.3. Metal Mixtures Mixtures of m e tals (including m e talloi ds and other contaminants) are commonly encountered in the natural environment as a result of anthropogenic inputs and should be considered b y the risk assessor for all assessm en ts. Meta l inte ractions, ac c o rding to Ca lam ari and Alabaster (1980), occur at three levels:

1. Chemical in tera ction s with oth er con stitu ents in the m edia, 2. Inter actions with th e physiological processes of the organism , and 3. Interactions at the si te of toxic action. The joint action of m etal mi xtures m a y be e xpressed in different ways, such as increasing or decr easin g the tox i city relative to that p r edic ted for individual components. As a result, the toxicity of m etal mi xtures has im portant consequences for m etals risk assessm ents. However, predicting the toxicity of m e tal mi xtures has proven to be a difficult challenge in ecotoxicology. Much of the dif f i culty ininte rpre ting the availab l e inf o rm ation on the tox i c ef f

ects of
m e tal mi xtures is due to differences in the bioava ilability of m e tals (and m easures used to define the bioav a ilable fraction ) that occur across mi xture studies. As disc ussed in Section 3, the bioavailability of m e tals depe nds on a suite of factors affec ting their speciation, comp lexation with lig ands , and inte rac tion with b i o l ogica l sys t e m s. Nevertheless, the risk assesso r needs som e m easure of the bioavailable m etal fraction in the exposure m e dia to accurately predict the effects of m e tals mi xtures (Sauv é et al., 199 8; W e ltje, 1998; Posthu m aet al., 1997). Besides bioavailability issues, the join t action of m e tal mixtures candepend on the overall mi xture concentrations and the relative pr opo rtion of the constituent m etals , as has been seen in aquatic studies (Norwood et al., 2003; Mowat and Bundy, 2002; Fargašová, 2001; Sharm aet al. 1999). The two m o st comm on c l asses of m o dels us ed to predic t mi xture tox i city are th e Concentration Addition and Effects Addition m o dels. These models have been used to classify the com bined effects of chemical mi xtures as being less than additive (i.e., when the observed ef

f

ect is les s than the m odel pred ictio n ), str i ctly a dditive (i.e., m a tching model predictions), and more than additive (i.e., when the observed effect is gr eater than m odel pr edictions; Norwood et al., 2003). Both m odels use m e tal concentrations in m edia to generate concentration-response curves for individual m e tals, and these data ar e th en used to g enerate spec if ic critical concentrations for mi xture m odels. In the Concen tration Add ition m odel, all m etals ina mi xture are added together to predict toxicity; differing potencies are taken into account by converting
6-14

### chemical concentration s to an equito xic dose (e .g., Toxic Units (TUs) or T oxicity Equivalence Factors (TEFs), which converts a ll m etals to one m etal concentr ation )

. Concentration Addition is used oftenwhen the co nstitu ents are known or assum ed to act through the sam e or similar MOA. However, th e ris k assesso r s hould use caution when applying the Concentration Addition model to mi xtures containing m a ny m e tal consti tuents (particularl y thos e well be low toxic levels) because of the potential for an upward bias in predicted mi xture t oxicity (Newm a n et al., 2004). In th e Effects Addition m odel, differing potencies are ignored, andthe effect of each
m etal’s conc entration is com b ined to predic t mi xture toxicity. The Effect Addition m o del is often used when constituents act independently (i.e., differentmodes of action). Only the Concentration Addition model allow s detec tion o f toxicity thatis m ore than additive. Thus, a key issue in applying either the Concentration Addition or Effect s Addition m odel is to define the natu re of the m etals’ joint action ( i .e., inde pendent or similar m ode of a c tion). The risk assessor can use inform a tion on the MOA, capacities to act as analogues for other m e tals, essentialities and ligand binding tend encies to m a ke this decision. Risk assessors should keep in mi nd, however, th at toxicities of certain m etal elem ents are associated w ith deficiencies of others. For exam ple, increased Zn, Cu, and Ni toxicities can be associated w ith 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 th is behavior can affect the uptake of m e tal elem ents (Marschner, 1998). Similar inte ractions occur in wildlife; for example, Cu toxicity can be a result of Modeficiency and vice versa (McDow ell, 2003; NAS/NRC, 1994a, 1980). It is po ssib l e that recep to r binding m odels (e .g., FIAM) m a y be expandedin the future to include mi xtures. In theory, if two m e tals comp et e for binding to the sam e site of toxic action on an organism , it should be possible to m odel the total m e tal bound to that site and, hence, to predic t m etal toxicity u sing a m echanistic re ceptor binding approach inanEffects Addition model. Alte rnatively, if two m e tals do not compe t e for the sam e binding site on the organism , then these models m a y provide m ore reliable estim ates of individual m e tal bioavailability, and these es tim ates can then be com b ined in m o re accurate Effects Additio n models. However, at present, thes e possibilitie s rem ain the o retical. Ad dition ally, th is possib i lity , while im proving the ability to ass ess the ef f

ec t s of m etal mi xtures, does not include tem poral aspects (i.e., “tim e-toresponse” versus concentration). From the preceding d i scussion, it sh ould be clear that the accurate pred iction of join t toxic ity of m etal mi xtures to terr es tria l organism s re m a ins a significant challenge.

6.3.4. Critical Body Residues Critical body residues (CBRs) are internal concen trations of chemicals that are correlated with the onset of a toxic res ponse (Conder et al., 2002; Lanno et al., 1998). CBRs can be based 6-15

### on whole-body residues (see below fo r discussion of this approachin soil invertebrates) or concentratio ns in specific tissu es. T he risk asse ssor m a y choose to use C BRs instead of dietary TRVs to reduce uncertainties b ecaus e they accou n t for site-sp ecific bio a v a ilab i lity an d multipathway issues (Van Straalen, 1 996; Van Wensem et al., 1994). Unfortunately, there are
m a jor data gaps in available CBRs fo r m a ny species—m etal co m b inations. Risk assesso rs can use tissue-specific critic al loa d s f o r som e m etals that h ave been established f o r several species of vertebrate wildlife, including Pb in liver, Cd in kidney, Hg in brains, and Se in eggs. See Beye r et al. (1996) f or these figures. Only a few CBRs have been developed in soil invertebrates for m e tals (C onder et al., 2002; Cromm e ntuijn et al., 1997, 1994; and Smit, 1997 for Cd and Zn). For plants, the use of a tissue residue (CBR) approach is another m e thod that risk
assessors mi ght use to addre ss m e tal toxicity issues, based on the concept that a m e tal concentratio n mu st reacha thresho l d value in the organism or at the targ et site b efore effects begin to occur (McCarty and Mackay, 1993; Lan no and McCarty, 1997). For essential elem ents in plan ts, deficiency/sufficien cy concentrations in foliage have been developed. However, the relationship between toxicity and tissue residues is complex a nd varies depending on tissue type (roots vs. shoots), plant species, and m e tal and th ere is little to no inf o rmation av ailable. 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 as s essor can estim ate TRVs (i.e ., toxic thresholds) for plants and soil invertebrates from laboratory tests where m etals are mi xed with sta ndard soils (Fairbrother et al., 2002). Variability am ong soil toxicity test results is due in part to the influence of soil properties on bioavailablity of m e tals (e.g., pH, organic m a tter and CEC). Additionally, acclim ation and adaptation of test organism s can furt her co mp licate test results and aging and other physical/chemi cal processes that affect m e tal sp eciation and uptake are not represented. Because incorporation of sparingly solubl e substances, such as m a ny envir onm ental f o rm s of m e tals, into the soil m atrix is dif f i cult, tes t s gene r ally ar e con ducted us ing soluble m etal sa lts with the addition of organism to the test m a trix imm e di ately after mi xing. The risk assessor should be aware of how all these factor s influence the test outcom e and subsequent TRV derivation. There is a large body of literature on toxicity of m e tals to soil organism s (e.g., van Straalen and Løkke, 1997), although often the objec tives were to understand processes rather than to develop defensible toxicity thresholds. T h e challenge for the risk assesso r, therefore, lies in how to use these data, taking into account th e test-to-test variabil ity in soil chemistry param e ters, and how to develop a technically defensible m eans of ex trapolating toxicity responses across soil type—in other words, how to adjust the toxicity threshold values for bioavailability differences in te st conditions. One approach to addressing variability in soil 6-16

### toxicity tests is to norm a lize test results by dividing the LC 50 

### (or, m o re generally, the E C x 

### ) by percent organic m a tter (Lock and Janssen, 2001)

. This approach is based on observed correlations between the LC 50 

### of Cu to earthworm s and soil organic m a tter content (Lock and Janssen, 2001). More recently, CEC has been show n to be the m ost i mpo rtant factor modifying Zn bioavailability in so ils for both in verteb rates and plants. B ecause CEC is a function , at least in part, of soil pH, norm a lization using this parameter should be done only am ong soi l s of similar pH ranges. However, comparison of field data wi th labo rato ry toxicity response inform ation still is best acco mplished by m easuring m etals in soil pore water from field assessm ents and
com p aring such data to s p iked labo ratory soils. R i sk assessors can us e the guidance docum e nt developed f or establishing Ecological Soil Sc reening Levels or Eco SSL to judge the applic ability of liter atur e studies to plant or so il in verteb rate to xicity thr esh old dete rminations. Eco SSLs have been developed for several m e tals, and the risk assessor should refer to these for national or regional assessm ents and for sc reen in g level, site-specific assessm ents.

6.3.6. Wild life Toxicity Toxicity in wildlife from m e tals exposures is generally poorly understood and is rarely quantified in field settings. A few notable exceptions are tho s e m echanism s described in avian waterfowl exposure to S e (Adam s et al., 2003) , exposure of waterfowl to Pb-contaminated sedim e nts (Henny et al., 2000; Beyer et al., 1998 ; Blus et al., 1991), and white-tailed ptarmi gan exposure to Cd in vegetation (Lariso n et al., 200 0). Most m e tals express multiorgan toxicity, resulting in a decrease in overall vigor, as opposed to well-definedm echanism s of ac tion docum ented from organic xenobiotics such as pestic ides. Typically, toxicological data used to assess the ri sk of m any m etals to wildlif e ar e deri ved from laboratory specie s such as rats, mi ce, or dom estic livestock species (e.g., cattle and ch ickens) exposed to soluble m e tal salts. Risk
assesso rs will need to ex trapo l ate the results of su ch tes t s to s p ecies of in terest b ecaus e of the paucity of data on the toxicity of m e tals to wildlife. However, risk assessors should approach this carefully due to the large am ount of uncertainty that could be introduced into the risk
assessm ent process (Suter, 1993). Laboratory and dom estic species m a y be m ore orless sen sitiv e to chemicals than are the selected wildlife species. Toxicolog i cal responses vary am ong species b ecause of many physiological factors that influe nce the toxicokinetics (absorpti on, distribution, and elimination) and toxicodynami cs (relative pote ncy) of m e tals after exposure has occurred. For example, differences in digestive tract physiology, renal ex cretion rates, and egg production influence the toxicokinetics of m e tals. The abil ity of som e species to m o re ra pidly produce protective proteins such as m e tallothionein after exposure to m e ta ls is a toxicodym a mi c features leading to interspecific extrapolation uncertainty. Thus, ri sk assessors should not extrapolate data from
m a mmal studies to birds, and should be aware that extrapolation of data from rats (simple, monogastric digestive physiology) to rumi na nts introduces m ore uncertainty than does 6-17

### extrapolation from rats to canids, and so on. In the c ase o f m etals, which som e species are ab le to regulate or s t ore in their tissues without experiencing toxic effects (i.e., biota-specific detoxifi cation), extrapolations between species used to assess bioaccumu lation and toxicity can be especially p r oblem atic. These difficulties in interspecific extrapolations are not unique to metals risk assessment except when dealing with essential elem ents. A review of potential extrapolation methodologies can be found in Kapustka et al. (2004). Currently, the best sources of inform ation for the risk assessor on w ildlife m e tal toxicity thres holds are NAS/NRC (1994a, 1980), McDowell (2003), and the docum e ntation s upporting developm ent of Eco SSLs values. The Eco SSL docum ent also includes an approach for screening studies fo r accep tability f o r use in derivation of toxicity thres holds for risk assessm ents, which can then be used for deriving sitespecific TR Vs for the most applicable endpoints. Risk assessors should apply uncertainty factors for extrapolation of data to species in a di fferent taxonomic category (e.g., genus, family or class) with caution and include a discussion of uncertainty inthe risk characterization

. Summ aries for som e m e t a ls are available in Beyer et al. (1996) and Fair brother et al. (1996). Risk Chara c teriza tion Have the qu a litative asse ssment, qua ntitative asse ssm ent, and key unce r tainties reg a rding metals be en pre s ente d in accordan ce with EPA guidelines ? Do conc lus ions fully reflec trisk s in relation to ambient con c entration s, essentiality of metals, chemi c al spe c iation, an d informatio n on variability in speci e s sensiti v ity? Have a s sump tions an d un certaintie s bee n documente d adeq uately? Have availabl e data on me cha n ism s of action and metal int e ra ction s bee n fully explored in developin g th e quantitative assessme nt in acc o rdanc e with EPA Guidanc e on Mix t ures Risk Asse ssment?



### 6.4. RISK CHARACTERI Z ATION Risk Characterization is the final phase of the ris k assessm ent proces s, in which inform ation from hazard characterization; dos e-response assessm e nt and exposure assessm ent are jointly considered to dete rmine the actual likelihood of risk to exposed populations (U.S. EPA, 2000c, 1998a). The characterization also sh ould discuss the uncertainties in the exposure and effects assessm ents, and the level of confidence in th e ove rall dete rmination of risk. At the
sam e tim e, Risk Charac teri zation is the f i rs t pha se in the risk m a nagem e nt process, in which inform ation from the characterizationis integratedinto the consequences of rule-m aking or risk
m a nage m e nt , such as consideration of cost, a lte rnative so lution s, politica l c onsideration s, community intera ction s. Each Risk Characterization should include three com ponents: a quali tative summary of each sectionof the risk assessm ent, a num erical risk estim ate, and a desc ription of assu mptions and uncertainties. These descriptions of variabi lity and uncertainty ar e particu l arly importan tf o r m e tals risk assessm ents given all the component s and challenges discussed in this Fr am e w ork docum ent. These are in addition to the variability and uncertainties that are inherent in all risk 6-18

### assessm ents (e.g., species to spec ies toxicity extrapolations). Because inf orm ation, knowledge, and tools are lacking for m a ny of the m e tal-speci f i c uncertainties, risk assessors should be particularly dilig ent in d o cum enting whether th es e m a y result in an over- or under-estim ation of risk (i.e., result in a conservative risk estim ate or not). It is likely thatsite -specif i c ris k
assessm ents will have fe wer uncertainties than re gional o r national s cale a ssessm ents because risk assesso rs have access to lo cal data on key issues such as specif i c m etal spec ies, r elative bioavailability, or background m e tal levels. For national or regional asse ssm e nts, selection of ranges or specific num bers for these values wi ll depend upon the degree of conservatism desired by the risk assessor and, ther efore, should be clearly docum ented during the Risk Characterization phase.

6-19

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### R-24

### APPE NDI X A: FRAMEWORK FOR MET A LS RI SK ASSES S MENT CONTRIBUTORS AND RE VIEWERS

CONTRIB U TORS AND REVIE W ERS Environmental Chemistry: Ki m Anderson, Oregon State University, Corvallis, OR Rufus Chaney, U.S. De partm e nt of Agriculture, Beltsville, MD Paul Chrostowski, CPF Associates, Inc., Takom aPark, MD Miriam L. Diam ond, Uni versity of Toronto, Toronto, ON, Ca nada Donald Langmu ir, Hydrochem System s Corp., Colorado School of Mines, Golden, CO Deborah Luecken, U.S. EPA, Office of Researchand Developm ent, National Exposure Research Laborato ry, Research Triangle Park, NC Michae l Ne wm an, College of W illia m and Mary, Glouceste r 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 Univer sity of New York, Stony Brook, NY Gerry Henningsen, H&H Scientific Services, Centennial, CO Rom an Lanno, Ohio State University, Colum bus, OH Jim McGeer, Natural Resources Canada, Ottawa, ON, Canada Exposure: Gary Diam ond, Syracuse Research C orp., Syracuse, NY Charles A. Menzie, Menzie-Cura & Associates, Inc., Chelm s ford, MA Jacqueline Moya, U.S. EPA, Office of Research and Deve lo pm ent, National Cente rf o r Environm ental Assessment, W a shington, DC Michae l Ne wm an, The College of W illiam and Mary, Glou ceste r Point, VA Jerom e Nriagu, University of Michigan, Ann Arbor, MI
Human He alth Effec t s : Ed Bender, U.S. EPA, Office of Research and Developm ent, Office of Science Policy, Washington, DC (Retired) Michael Beringer, U.S. EPA, Region VII, Kansas City, KS Harlal Choudhury, U.S. EPA, Office of Res earch and Developm ent, National Center for Environm ental Assessment, Cincinnati, OH Do mi nic Di Toro, University of Delaware, Newark, DE Mari Hirtzel Golub, University of California, Davis, CA Robert Goyer, University of W e stern Ontario, N a tional Research Council Board Member/Chair Mike Hughes, U.S. EPA, Office of Research a nd Developm ent, National Health and Ecological Effects Research Laboratory, W a shington, DC Richard Hertzberg, U.S. EPA, Office of Research and Deve lo pm ent, National Cente rf o r Environm ental Assessment, Atlanta, GA (Retired) Elaina Kenyon, U.S. EPA, Office of Research and Development, National Health and Ecological Effects Research Laboratory, W a shington, DC A-1

### Fran Krem er, U.S. EPA, Office of Research and Developm ent, Nati onal Risk Managem e nt Research Laboratory, W a shington, D C Deirdre Murphy, U.S. EPA, Office of Air and Radiation, Research Triangle Park, NC Ecological Effects: W illiam Clem ents, Colorado Stat e Univers ity, Ft. Collins, CO Heidi Bell, U.S. EPA, Offi ce of Water, W a shington, DC Peter Chapm a n, EVS Environm ent Consultants, N. Vancouver, BC, Canada Dale Hoff, U.S. EPA, Region VIII, Denver, CO Larry Kapustka, Ecological Planning and Toxicology, Inc., Corvallis, OR Michae l Ne wm an, The College of W illiam and Mary, Glou ceste r Point, VA Paul Paquin, HydroQual, Inc., Mahw ah, NJ Cindy Roberts, U.S. EPA, Offi ce of Water, W a s hington, DC Mark Sprenger, U.S. EPA, Office of Solid W a ste and Em erge ncy Response, W a shington, DC Dan W a ll, U.S. Department of the Interior, U. S. Fish and W i ldlife Se rvice, W a shington, DC Linda Ziccardi, Exponent, Boulder, CO
METALS ISSUE P A P E R AUTH ORS 1 

### Environmental Chemistry of Metals: Donald Langmu ir, Hydrochem System s Corp., Colorado School of Mines, Golden, CO Paul Chrostowski, CPF Associates, Inc., Takom aPark, MD Rufus Chaney, U.S. De partm e nt of Agriculture, Beltsville, MD Bernard Vigneault, CANMET Mining and Minera l Scien ces Laborato r ies /Natura l Res ources Canada, Ottawa, ON Metal Expo sure Assessment: Michae l Ne wm an, The College of W illiam and Mary/VIM S, Glouceste r Point, VA Gary Diam ond, Syracuse Research C orp., Syracuse, NY Charles A. Menzie, Menzie-Cura & Associates, Chelm s ford, MA Jerom e Nriagu, University of Michigan, Ann Arbor, MI Ecological Effects o f Metals : Lawrence Kapustka, Ecological Planning and Toxicology, Inc., Corvallis, OR W illiam Clem ents, Colorado Stat e Univers ity, Fo rt Collins, C O Linda Ziccardi, Exponent, Boulder, CO Paul Paquin, HydroQual, Inc., Mahw ah, NJ Mark Sprenger, U.S. EPA, Edison, NJ Daniel W a ll, U.S. Fish and W ildlife Service, Denver, CO
Human He alth Effec t s of Meta l s: Robert Goyer, University of W e stern Ontario, O N 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, ResearchTriangle Park, NC (contributor) Marc Stif elm a n, U.S. EPA, Seattle, W A(contrib u tor)

A-2

### Bioavailability and Bioaccumulation of Metals: John Drexler, University of Colorado, Boulder, CO Nicholas Fisher, State Univer sity of New York, Stony Brook, NY Gerry Henningsen, H&H Scientific Services, Centennial, CO Rom an Lanno, Ohio State University, Colum bus, OH Jim McGeer, Natural Resources Canada, Ottawa, ON Keith Sappington, U.S. EPA, Office of Resear ch and Developm ent, National Center for Environm ental Assessment, W a shington, DC Michael Beringer, U.S. EPA, Region VII, Kansas City, KS (contributor) PEER CONSUL TATI ON WORKSHOP PARTICIPANT S Breakout Group 1 (Overall Framew ork): Pat Doyle, E nvironm ent Canada, Ottawa, ON (breakout group lead) Charles A. Menzie, Menzie-Cura & Associ ates, Chelm s ford, MA ( W orkshop Chairman) Bruce K. Hope, Oregon Departm e nt of Environmental Quality, Portland, OR Breakout Group 2 (Environmental Chemistry): Rufus Chaney, U.S. De partm e nt of Agricu lture, Beltsville, MD (breakout group lead) Donald Langmu ir, Hydrochem System s Corp., Colorado School of Mines, Golden, CO Mark Barnett, Departm e nt of Civ il E ngineering, Auburn University, AL Rom an Lanno, Ohio State University, Colum bus, OH Jim Ryan, U.S. EPA, Office of Research a nd Developm ent, National Risk Managem e nt Research Laboratory, C i ncinnati, OH Breakout Group 3 (Terrestrial Exp osure/Effects): John Drexler, University of Colora do, Boulder, CO (breakout group lead) Lawrence Kapustka, Ecological Planning and Toxicology, Inc., Corvallis, OR Bradley Sa mple, CH2M Hill, Sac r am ento, CA Beverly Hale, University of Guelph, Guelph, ON Breakout Group 4 (Aquatic Exposure/Effects): Keith Sappington, U.S. EPA, Office of Resear ch and Developm ent, National Center for Environm ental Assessment, W a shi ngton, DC (breakout group lead) David Mount, U.S. EPA, Office of Researchand Developm ent, National Health and Environm ental Effects L aboratory, Duluth, MN Jim McGeer, Natural Resources Canada, Ottawa, ON Steve Klaine, Clem son University, Clem son, SC Paul Paquin, HydroQual, Inc., Mahw ah, NJ Anne Smith-Reiser, Analytical Services Inc., Jessup, MD Breakout Group 5 (Human Exposure/Hea l th Effects ) : Steve Devito, U.S. EPA, Office of Environm en tal Inform ation, W a shington, DC (breakout group lead) Mari Golub, University of California, Davis, CA Herm an Gibb, Sciences Inte rnational, Alexandria, VA Rosalind A. Schoof, Integral Cons ulting, Inc., Mercer Island, W A Deirdre Murphy, U.S. EPA, Office of Research and Developm ent, Office of Radiation, Office of Air Quality Planning and Standard s, Research Triangle Park, NC A-3

### ___________________ Alexander McBride, U.S. EPA, Office of So lid Waste and Em ergency Response, W a shington, DC Margaret MacDonell, U.S. Departm e nt of En ergy, Argonne National Laboratory, Argonne, IL Bob Jones, U.S. EPA, Office of Prevention, Pe sticides, and Toxic Substances, W a shington, DC Richard Troast, U.S. EPA, Office of Solid W a ste and Em erge ncy Response, W a shington, DC

1 

### Inform ation from the m etals issue papers wa s used in the developm ent of this fram e w ork

A-4PRESOR TED ST ANDARD POST AGE & FEES P AID EP A PERMIT NO. G-35

O f fi c e o f R e s e a r c h a n d D e v e l o p m e n t ( 8 1 0 1 R ) W a s h i n g t o n , D C 2 0 4 6 0 O f fi c i a l B u s i n e s s P e n a l t y f o r P r i v a t e U s e $ 3 0 0