October 2006 EPA/600/R-05/144bF 

# Air Quality Criteria for Lead

Volume II of IIEPA/600/R-05/144bF October 2006



# Air Quality Criteria for Lead

Volume II

National Center for Environmental Assessment-RTP Division Office of Research and Development U.S. Environmental Protection Agency Research Triangle Park, NCII-iii PREFACE

National Ambient Air Quality Standards (NAAQS ) are promulgated by the United States Environmental Protection Agency (EPA) to meet requirements set forth in Sections 108 and 109

of the U.S. Clean Air Act. Those two Clean Air Act sections require the EPA Administrator (1) to list widespread air pollutants that reasonabl y may be expected to endanger public health or welfare; (2) to issue air qualit y criteria for them that assess the latest available scientific

information on nature and effects of ambient e xposure to them; (3) to set “primary” NAAQS to protect human health with ade quate margin of safety and to set “secondary” NAAQS to protect against welfare effects (e.g., effects on vegeta tion, ecosystems, visibility, climate, manmade materials, etc); and (5) to peri odically review and revise, as appropriate, the criteria and NAAQS for a given listed pollutant or class of pollutants.

Lead was first listed in the mid-1970’s as a “criteria air pollutant” requiring NAAQS regulation. The scientific inform ation pertinent to Pb NAAQS de velopment available at the time

was assessed in the EPA document Air Quality Criteria for Lead ; published in 1977. Based on

the scientific assessments contained in that 1977 lead air quality criteria document (1977 Lead AQCD), EPA established a 1.5 μg/m 3 (maximum quarterly calendar average) Pb NAAQS in

1978. To meet Clean Air Act requirements noted above for periodic review of criteria and NAAQS, new scientific information published si nce the 1977 Lead AQCD was later assessed in a revised Lead AQCD and Addendum published in 1986 and in a Supplement to the 1986 AQCD/Addendum published by EPA in 1990. A 1990 L ead Staff Paper, prepared by EPA’s Office of Air Quality Planning and Standards (OPQPS), drew upon key findings and conclusionsfrom the 1986 Lead AQCD/Addendum and 1990 Supplement (as well as other OAQPSsponsored lead exposure/risk analyses) in posing options for the EPA Administrator to consider with regard to possible revision of the Pb NAAQS . However, EPA chose not to revise the Pb NAAQS at that time. Rather , as part of implementing a broad 1991 U.S. EPA Strategy for Reducing Lead Exposure, the Agency focused primarily on regulatory and remedial clean-up efforts to reduce Pb exposure from a variety of non-air sources that posed more extensive public health risks, as well as other actions to reduce air emissions.II-iv The purpose of this revised Lead AQCD is to critically assess the latest scientific

information that has become available sin ce the literature asse ssed in the 1986 Lead AQCD/Addendum and 1990 Supplement, with the main focus being on pertinent new information useful in evaluating health and enviro nmental effects of ambien t air lead exposures.

This includes discussion in this document of information rega rding: the nature, sources, distribution, measurement, and c oncentrations of lead in the environment; multimedia lead exposure (via air, food, water, et c.) and biokinetic modeling of c ontributions of such exposures to concentrations of lead in brain, kidney, a nd other tissues (e.g., blood and bone concentrations, as key indices of lead exposure).; characterization of l ead health effects and associated exposureresponse relationships; and delineation of environmen tal (ecological) effects of lead. This final version of the revised Lead AQCD mainly assesses pertinent literature published or accepted for publication through December 2005. The First External Review Draft (dated December 2005) of the revised Lead AQCD underwent public comment and was reviewed by the Clean Air Scientific Advisory Committee (CASAC) at a public meeting held in Durham , NC on February 28-March 1, 2006. The public

comments and CASAC recommendations receiv ed were taken into account in making appropriate revisions and incorporating them in to a Second External Review Draft (dated May, 2006) which was released for further public co mment and CASAC review at a public meeting held June 28-29, 2006. In addition, still further revised drafts of the Integrative Synthesis chapter and the Executive Summary were then issued and discussed during an August 15, 2006

CASAC teleconference call. Public commen ts and CASAC advice received on these latter materials, as well as Second External Review Draf t materials, were taken into account in making and incorporating further revisions into this fi nal version of this Lead AQCD, which is being issued to meet an October 1, 2006 court-ordered d eadline. Evaluations c ontained in the present document provide inputs to an associated Lead Staff Paper prepared by EPA’s Office of Air Quality Planning and Standards (OAQPS), which poses options for consideration by the EPA Administrator with regard to proposal and, ul timately, promulgation of decisions on potential retention or revision, as appropriate, of the current Pb NAAQS. Preparation of this document has been coordi nated by staff of EPA’s National Center for Environmental Assessment in Research Triangl e Park (NCEA-RTP). NCEA-RTP scientific staff, together with experts from academia, contri buted to writing of docum ent chapters. EarlierII-v drafts of document materials were reviewed by scientists from other EPA units and by non-EPA experts in several public peer consultation workshops held by EPA in July/August 2005. NCEA acknowledges the valuab le contributions provided by authors, contributors, and reviewers and the diligence of its staff and contractors in the prep aration of this document. The constructive comments provided by public comment ers and CASAC that served as valuable inputs contributing to improved scientific and editorial quality of the document are also acknowledged by NCEA.

DISCLAIMER

Mention of trade names or commercial produc ts in this document does not constitute

endorsement or recommendation for use.

.II-vi 

### Air Quality Criteria for Lead (Second External Review Draft)

VOLUME I

EXECUTIVE SUMMARY .........................................................................................................E-1

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

2. CHEMISTRY, SOURCES, AN D TRANSPORT OF LEAD.......................................... 2-1

3. ROUTES OF HUMAN EXPOSURE TO LEAD AND OBSERVED ENVIRONMENTAL CONCENTRATIONS.................................................................. 3-1

4. TOXICOKINETICS, BIOLOGICAL MARKERS, AND MODELS OF LEAD BURDEN IN HUMANS.................................................................................................. 4-1

5. TOXICOLOGICAL EFFECTS OF LEAD IN LABORATORY ANIMALS AND IN VITRO TEST SYSTEMS................................................................................. 5-1

6. EPIDEMIOLOGIC STUDIES OF HUMAN HEALTH EFFECTS ASSOCIATED WITH LEAD EXPOSURE.................................................................... 6-1

7. ENVIRONMENTAL EFFECTS OF LEAD ................................................................... 7-1

8. INTEGRATIVE SYNTHESIS OF LEAD EXPOSURE/HEALTH EFFECTS INFORMATION............................................................................................ 8-1



### VOLUME II

CHAPTER 4 ANNEX (TOXICOKINETICS , BIOLOGICAL MARKERS, AND MODELS OF LEAD BURDEN IN HUMANS)..................................................... AX4-1

CHAPTER 5 ANNEX (TOXICOLOGICAL EFFECTS OF LEAD IN LABORATORY ANIMALS AND IN VITRO TEST SYSTEMS)................ AX5-1

CHAPTER 6 ANNEX (EPIDEMIOLOGIC STUDIES OF HUMAN HEALTH EFFECTS ASSOCIATED WITH LEAD EXPOSURE)................................. AX6-1

CHAPTER 7 ANNEX (ENVIRONMENTA L EFFECTS OF LEAD).................................. AX7-1II-vii Table of Contents

Page

PREFACE........................................................................................................................ ..............iii DISCLAIMER ..................................................................................................................... ........... v List of Tables ................................................................................................................. ................ xi List of Figures................................................................................................................ ..............xiii Authors, Contributors, and Reviewers.......................................................................................... x v U.S. Environmental Protection Agency Project Team for Development of Air Quality Criteria for Lead.......................................................................................... .... xxi U.S. Environmental Protection Agency Science Advisory Board (SAB) Staff Office Clean Air Scientific Advisory Committee (CASAC).......................................xxiii Abbreviations and Acronyms ..................................................................................................... xxv AX7. CHAPTER 7 ANNEX (ENVIRONMEN TAL EFFECTS OF LEAD).................... AX7-1 AX7.1 TERRESTRIAL ECOSYSTEMS............................................................ AX7-1 AX7.1.1 Methodologies Used in Terrestrial Ecosystems Research............................................................................... AX7-1 AX7.1.1.1 Lead Isotopes and Apportionment................. AX7-1 AX7.1.1.2 Speciation in Assessing Lead Bioavailability in the Terrestrial Environment................................................... AX7-3 AX7.1.1.3 Tools for Bulk Lead Quantification and Speciation................................................ AX7-9 AX7.1.1.4 Biotic Ligand Model.................................... AX7-18 AX7.1.1.5 Soil Amendments......................................... AX7-19 AX7.1.2 Distribution of Atmosp herically Delivered Lead in Terrestrial Ecosystems....................................................... AX7-22 AX7.1.2.1 Speciation of Atmo spherically-Delivered Lead in Terrestrial Ecosystems.................... AX7-24 AX7.1.2.2 Tracing the Fate of AtmosphericallyDelivered Lead in Terrestrial Ecosystems................................................... AX7-31 AX7.1.2.3 Inputs/Outputs ofAtmosphericallyDelivered Lead in Terrestrial Ecosystems................................................... AX7-34 AX7.1.2.4 Resistance Mechanisms ............................... AX7-44 AX7.1.2.5 Physiological Effects of Lead ...................... AX7-46 AX7.1.2.6 Factors that Modify Organism Response..... AX7-48 AX7.1.2.7 Summary...................................................... AX7-55 AX7.1.3 Exposure-Response of Terrestrial Species......................... AX7-57 AX7.1.3.1 Summary of Conclusions from the 1986 Lead Criteria Document...................... AX7-59 AX7.1.3.2 Recent Studies on the Effects of Lead on Primary Producers................................... AX7-61II-viii Table of Contents (cont’d)

Page

AX7.1.3.3 Recent Studies on the Effects of Lead on Consumers............................................... AX7-64 AX7.1.3.4 Recent Studies on the Effects of Lead on Decomposers........................................... AX7-81 AX7.1.3.5 Summary...................................................... AX7-86 AX7.1.4 Effects of Lead on Natu ral Terrestrial Ecosystems ........... AX7-88 AX7.1.4.1 Effects of Terrestrial Ecosystem Stresses on Lead Cycling........................................... AX7-89 AX7.1.4.2 Effects of Lead Exposure on Natural Ecosystem Structure and Function............... AX7-95 AX7.1.4.3 Effects of Lead on Energy Flows and Biogeochemical Cycling.............................. AX7-99 AX7.1.4.4 Summary.................................................... AX7-105 AX7.2 AQUATIC ECOSYSTEMS................................................................. AX7-106 AX7.2.1 Methodologies Used in Aquatic Ecosystem Research..... AX7-106 AX7.2.1.1 Analytical Methods.................................... AX7-107 AX7.2.1.2 Ambient Water Quality Criteria: Development.............................................. AX7-108 AX7.2.1.3 Ambient Water Quality Criteria: Bioavailability Issues................................. AX7-111 AX7.2.1.4 Sediment Quality Criteria: Development and Bioavaila bility Issues.... AX7-113 AX7.2.1.5 Metal Mixtures........................................... AX7-116 AX7.2.1.6 Background Lead....................................... AX7-117 AX7.2.2 Distribution of Lead in Aquatic Ecosystems ................... AX7-117 AX7.2.2.1 Speciation of Lead in Aquatic Ecosystems................................................. AX7-118 AX7.2.2.2 Spatial Distribution of Lead in Aquatic Ecosystems................................... AX7-122 AX7.2.2.3 Tracing the Fate and Transport of Lead in Aquatic Ecosystems...................... AX7-142 AX7.2.2.4 Summary.................................................... AX7-145 AX7.2.3 Aquatic Species Re sponse/Mode of Action..................... AX7-146 AX7.2.3.1 Lead Uptake............................................... AX7-146 AX7.2.3.2 Resistance Mechanisms ............................. AX7-152 AX7.2.3.3 Physiological Effects of Lead .................... AX7-159 AX7.2.3.4 Factors That Modify Organism Response to Lead ....................................... AX7-162 AX7.2.3.5 Factors Associated with Global Climate Change.......................................... AX7-174 AX7.2.3.6 Summary.................................................... AX7-174II-ix Table of Contents (cont’d)

Page

AX7.2.4 Exposure/Response of Aquatic Species........................... AX7-175 AX7.2.4.1 Summary of Conc lusions From the Previous Criteria Document....................... AX7-175 AX7.2.4.2 Recent Studies on Effects of Lead on Primary Producers...................................... AX7-177 AX7.2.4.3 Recent Studies on Effects of Lead on Consumers............................................. AX7-183 AX7.2.4.4 Recent Studies on Effects of Lead on Decomposers......................................... AX7-193 AX7.2.4.5 Summary.................................................... AX7-193 AX7.2.5 Effects of Lead on Na tural Aquatic Ecosystems ............. AX7-194 AX7.2.5.1 Case Study: Coeur d’Alene River Watershed .................................................. AX7-196 AX7.2.5.2 Biotic Condition......................................... AX7-198 AX7.2.5.3 Summary.................................................... AX7-209 REFERENCES .................................................................................................... AX7-211II-x List of Tables

Page

AX7-1.1.1 Relative Standard Deviation (R SD) for Lead Isotope Ratios on Selected Mass Spectrometers................................................................... AX7-2

AX7-1.1.2 National Institute of Standa rds and Technology Lead SRMs................ AX7-10

AX7-1.1.3 Characteristics for Direct Speciation Techniques.................................. AX7-17

AX7-1.2.1 Tissue Lead Levels in Birds Causing Effects ........................................ AX7-43

AX7-1.3.1 Plant Toxicity Data Used to Develop the Eco-SSL............................... AX7-62

AX7-1.3.2 Plant Toxicity Data Not Us ed to Develop the Eco-SSL........................ AX7-63

AX7-1.3.3 Avian Toxicity Data Used to Develop the Eco-SSL ............................. AX7-66

AX7-1.3.4 Mammalian Toxicity Data Us ed to Develop the Eco-SSL.................... AX7-72

AX7-1.3.5 Invertebrate Toxicity Data Used to Develop the Eco-SSL.................... AX7-82

AX7-1.3.6 Invertebrate Toxicity Data No t Used to Develop the Eco-SSL............. AX7-84

AX7-2.1.1 Common Analytical Methods for Measuring Lead in Water, Sediment, and Tissue .......................................................................... AX7-107

AX7-2.1.2 Development of Curre nt Acute Freshwater Cr iteria for Lead ............. AX7-109

AX7-2.1.3 Recommended Sediment Quality Guidelines for Lead ....................... AX7-115

AX7-2.2.1 NAWQA Land Use Categorie s and Natural/Ambient Classification........................................................................................ AX7-125

AX7-2.2.2 Summary Statistics of Ambient and Natural Levels of Dissolved Lead in Surface Water ......................................................................... AX7-126

AX7-2.2.3 Summary Statistics of Ambien t and Natural Levels of Total Lead in <63 μm Bulk Sediment........................................................... AX7-127

AX7-2.2.4 Summary Statistics of Ambient and Natural Levels of Lead in Whole Organism and Liver Tissues..................................................... AX7-136

AX7-2.2.5 Comparison of NCBP and NA WQA Ambient Lead Levels in Whole Organism Tissues ..................................................................... AX7-138II-xi List of Tables (cont’d)

Page

AX7-2.3.1 Bioconcentration Factors for Aquatic Plants ....................................... AX7-151

AX7-2.3.2 Bioconcentration Factors fo r Aquatic Invertebrates............................ AX7-151

AX7-2.4.1 Effects of Lead to Freshwater and Marine Invertebrates..................... AX7-185

AX7-2.4.2 Effects of Pb to Freshw ater and Marine Fish....................................... AX7-190

AX7-2.4.3 Nonlethal Effects in Amphibians......................................................... AX7-192

AX7-2.5.1 Ecological Attributed Studies by Maret et al. (2003) in the Coeur d’Alene Watershed.................................................................... AX7-197

AX7-2.5.2 Essential Ecological Attributes for Natural Aquatic Ecosystems Affected by Lead.................................................................................. AX7-200II-xii List of Figures

Number Page

AX7-1.1.1 Relationship of bioaccessibil ity (low, medium, high) versus speciation as shown with scanning electron micrographs of various Pb-bearing materials........................................................................ AX7-5

AX7-1.1.2 Variation of bioavailabil ity with particle size.............................................. AX7-7

AX7-1.1.3 Illustration of particle lability and bioavailability at two different sites with similar total Pb concentrations and Pb forms .............................. AX7-8

AX7-1.1.4 Scanning electron micrograph of a large native Pb particle showing outer ring of highly bioavailable Pb-chloride and Pb-oxide........................ AX7-8

AX7-1.1.5 Bulk lead versus single species modality................................................... AX7-12

AX7-1.2.1 Long-term (1982-1989) annual trends in lead concentrations (μg/L) in Lewes, Delaware precipitation .............................................................. AX7-35

AX7-1.3.1 Avian reproduction and growth toxicity data considered in development of the Eco-SSL ..................................................................... AX7-69

AX7-1.3.2 Mammalian reproduction and growth toxicity data considered in development of the Eco-SSL ..................................................................... AX7-81

AX7-2.2.1 Distribution of aqueous lead sp ecies as a function of pH based on a concentration of 1 μg Pb/L.................................................................... AX7-119

AX7-2.2.2 Lead speciation versus chloride content .................................................. AX7-121

AX7-2.2.3 Spatial distribution of natural and ambient surface water/sediment sites .......................................................................................................... AX7-128

AX7-2.2.4 Spatial distribution of natural and ambient liver tissue sample sites....... AX7-129

AX7-2.2.5 Spatial distribution of natural and ambient whole organism tissue sample sites .............................................................................................. AX7-130

AX7-2.2.6 Frequency distribution of ambi ent and natural levels of surface water dissolved lead (μg/L). .................................................................... AX7-131

AX7-2.2.7 Spatial distribution of dissolved l ead in surface water (N = 3445). ........ AX7-132II-xiii List of Figures (cont’d)

Number Page

AX7-2.2.8 Frequency distribution of ambi ent and natural levels of bulk sediment <63 μm total Pb (μg/g)............................................................. AX7-134

AX7-2.2.9 Spatial distribution of total lead in bulk sediment <63 um (N = 1466)................................................................................................ AX7-135

AX7-2.2.10 Frequency distribution of ambien t and natural levels of lead in liver tissue (μg/g dry weight)................................................................... AX7-137

AX7-2.2.11 Frequency distribution of ambien t and natural levels of lead in whole organism tissue (μg/g dry weight). ............................................... AX7-137

AX7-2.2.12 Spatial distribution of lead in liver tissues (N = 559). ............................. AX7-139

AX7-2.2.13 Spatial distribution of lead in w hole organism tissues (N = 332)............ AX7-140

AX7-2.2.14 Lead cycle in an aquatic ecosystem......................................................... AX7-143II-xiv Authors, Contributors, and Reviewers

CHAPTER 4 ANNEX (TOXICOKINETIC S, BIOLOGICAL MARKERS, AND MODELS OF LEAD BURDEN IN HUMANS)

Chapter Managers/Editors

Dr. James Brown—National Center for Environmental Assessment (B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Principal Authors

Dr. Brian Gulson—Graduate School of the Environment, Macquarie University Sydney, NSW 2109, Australia

Contributors and Reviewers

Dr. Lester D. Grant—National Center for Environmental Assessment (B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. John Vandenberg—National Center for Environmental Assessment (B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

CHAPTER 5 ANNEX (TOXICOLOGICAL EFFECTS OF LEAD IN HUMANS AND LABORATORY ANIMALS)

Chapter Managers/Editors

Dr. Anuradha Mudipalli—National Center for Environmental Assessment (B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Srikanth Nadadur—National Center for Environmental Assessment (B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Lori White—National Center for Environmental Assessment (B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711II-xv Authors, Contributors, and Reviewers (cont’d)

Principal Authors

Dr. Anuradha Mudipalli—National Center for Environmental Assessment (B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Stephen Lasley—Dept. of Biomedical and Ther apeutic Sciences, Univ. of Illinois College of Medicine, PO Box 1649, Peoria, IL 61656

Dr. Lori White—National Center for Environmental Assessment (B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Gary Diamond— Syracuse Research Cor poration, 8191 Cedar Street, Akron, NY 14001

Dr. N.D. Vaziri—Division of Nephrology and Hype rtension, University of California – Irvine Medical Center, 101, The City Drive, Bldg 53, Room #125. Orange, CA 92868

Dr. John Pierce Wise, Sr.—Maine Center fo r Toxicology and Environmental Health, Department of Applied Medical Sciences, 96 Falmouth Street, PO Box 9300, Portland, ME 04104-9300

Dr. Harvey C. Gonick—David Geffen School of Medicine, University of California at Los Angeles, 201 Tavistock Ave, Los Angeles, CA 90049

Dr. Gene E. Watson—University of Rochester Medical Center, Box 705, Rochester, NY 14642

Dr. Rodney Dietert—Institute for Comparativ e and Environmental Toxicology, College of Veterinary Medicine, Cornel l University, Ithaca, NY 14853

Contributors and Reviewers

Dr. Lester D. Grant—National Center for Environmental Assessment (B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Paul Reinhart—Nationa l Center for Environmental Assessment (B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Michael Davis—Nationa l Center for Environmental Assessment (B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. David A. Lawrence—Dept of Environmental and Clinical Immunology, Empire State Plaza P.O. Box 509, Albany, NY 12201II-xvi Authors, Contributors, and Reviewers (cont’d)

Contributors and Reviewers (cont’d)

Dr. Michael J. McCabe, Jr.—Dept of Environm ental Medicine, University of Rochester, 575 Elmwood Avenue, Rochester, NY 14642

Dr. Theodore I. Lidsky—New York State Inst itute for Basic Research, 1050 Forest RD, Staten Island, NY 10314

Dr. Mark H. Follansbee—Syracuse Research Corporation, 8191 Cedar St. Akron, NY 14001

Dr. William K. Boyes—National Health and En vironmental Effects Research Laboratory, U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Philip J. Bushnell—National Health and Environmental Effects Research Laboratory, U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Ms. Beth Hassett-Sipple—Office of Air Quality Planning and Standards, U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Zachary Pekar—Office of Air Quality Pla nning and Standards, U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. John Vandenberg—National Center for Environmental Assessment (B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

CHAPTER 6 ANNEX (EPIDEMIOLOGICA L STUDIES OF AMBIENT LEAD EXPOSURE EFFECTS)

Chapter Managers/Editors

Dr. Jee Young Kim—National Center for Environmental Assessment (B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Dennis Kotchmar—National Center for Environmental Assessment (B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. David Svendsgaard—National Center for Environmental Assessment (B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711II-xvii Authors, Contributors, and Reviewers (cont’d)

Principal Authors

Dr. David Bellinger—Children’s Hospital, Fa rley Basement, Box 127, 300 Longwood Avenue, Boston, MA 02115

Dr. Margit Bleecker—Center for Occupational and Environmental Neurology, 2 Hamill Road, Suite 225, Baltimore, MD 21210

Dr. Gary Diamond— Syracuse Resear ch Corporation, 8191 Cedar Street. Akron, NY 14001

Dr. Kim Dietrich—University of Cincinnati College of Medicine, 3223 Eden Avenue, Kettering Laboratory, Room G31, Cincinnati, OH 45267

Dr. Pam Factor-Litvak—Columbia University Mailman School of Public Health, 722 West 168th Street, Ro om 1614, New York, NY 10032

Dr. Vic Hasselblad—Duke University Medical Center, Durham, NC 27713

Dr. Stephen J. Rothenberg—CINVESTAV-IPN, Mé rida, Yucatán, México & National Institute of Public Health, Cuernavaca, Morelos, Mexico

Dr. Neal Simonsen—Louisiana State University Hea lth Sciences Center, School of Public Health & Stanley S Scott Cancer Center, 1600 Cana l Street, Suite 800, New Orleans, LA 70112

Dr. Kyle Steenland—Rollins School of Public Health, Emory University, 1518 Clifton Road, Room 268, Atlanta, GA 30322

Dr. Virginia Weaver—Johns Hopkins Bloomberg School of Public Health, 615 North Wolfe Street, Room 7041, Baltimore, MD 21205

Contributors and Reviewers

Dr. J. Michael Davis—National Center for Environmental Assessment (B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Lester D. Grant—National Center for Environmental Assessment (B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Kazuhiko Ito—Nelson Institute of Environmental Medicine, Ne w York University School of Medicine, Tuxedo, NY 10987II-xviii Authors, Contributors, and Reviewers (cont’d)

Contributors and Reviewers (cont’d)

Dr. Kathryn Mahaffey—Office of Preventi on, Pesticides and Toxic Substances, U.S. Environmental Protection Agency, Washington, DC 20460

Dr. Karen Martin—Office of Air Quality Pl anning and Standards, U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Zachary Pekar—Office of Air Quality Pla nning and Standards, U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Ms. Beth Hassett-Sipple—Office of Air Quality Planning and Standards, U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. John Vandenberg—National Center for Environmental Assessment (B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

CHAPTER 7 ANNEX (ENVIRONME NTAL EFFECTS OF LEAD)

Chapter Manager/Editor

Dr. Timothy Lewis—National Center for Environmental Assessment (B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Principle Authors

Dr. Ruth Hull—Cantox Environmental Inc., 190 0 Minnesota Court, Suite 130, Mississauga, Ontario, L5N 3C9 Canada

Dr. James Kaste—Department of Earth Scienc es, Dartmouth College, 352 Main Street, Hanover, NH 03755

Dr. John Drexler—Department of Geologica l Sciences, University of Colorado, 1216 Gillespie Drive, Boulder, CO 80305

Dr. Chris Johnson—Department of Civil and Envir onmental Engineering, Syracuse University, 365 Link Hall, Syracuse, NY 13244

Dr. William Stubblefield—Parametrix, Inc. 33972 Texas St. SW, Albany, OR 97321II-xix Authors, Contributors, and Reviewers (cont’d)

Principle Authors (cont’d)

Dr. Dwayne Moore—Cantox Environmental, Inc., 1550A Laperriere Avenue, Suite 103, Ottawa, Ontario, K1Z 7T2 Canada

Dr. David Mayfield—Parametrix, Inc., 411 108th Ave NE, Suite 1800, Bellevue, WA 98004

Dr. Barbara Southworth—Menzie-Cura & Associ ates, Inc., 8 Winchester Place, Suite 202, Winchester, MA 01890

Dr. Katherine Von Stackleberg—Menzie-Cura & Associates, Inc., 8 Winchester Place, Suite 202, Winchester, MA 01890

Contributors and Reviewers

Dr. Jerome Nriagu—Department of Environmenta l Health Sciences, 109 South Observatory, University of Michigan, Ann Arbor, MI 48109

Dr. Judith Weis—Department of Biol ogy, Rutgers University, Newark, NJ 07102

Dr. Sharon Harper—National Exposure Research Laboratory (D205-05), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Karen Bradham—National Research Exposur e Laboratory (D205-05), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Ginger Tennant—Office of Air Quality Plan ning and Standards, U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Ms. Gail Lacey—Office of Air Quality Planning a nd Standards, U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. John Vandenberg—National Center for Environmental Assessment (B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711II-xx U.S. Environmental Protection Agency Project Team

for Development of Air Quality Criteria for Lead

Executive Direction

Dr. Lester D. Grant (Director)—National Cent er for Environmental Assessment-RTP Division, (B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Scientific Staff

Dr. Lori White (Lead Team Leader)—Nati onal Center for Environmental Assessment (B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. James S. Brown—National Center fo r Environmental Assessment (B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Robert Elias—Nationa l Center for Environmental Assessment (B243-01), U.S. Environmental Protection Agency, Re search Triangle Park, NC 27711 (Retired)

Dr. Brooke Hemming—National Center for Environmental Assessment (B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Jee Young Kim—National Center for Environmental Assessment (B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Dennis Kotchmar—National Center for Environmental Assessment (B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Timothy Lewis—National Center for Environmental Assessment (B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Anuradha Muldipalli—National Center for Environmental Assessment (B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Srikanth Nadadur—National Center for Environmental Assessment (B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Paul Reinhart—Nationa l Center for Environmental Assessment (B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. Mary Ross— National Center for Environmental Assessment (B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Dr. David Svendsgaard—National Center for Environmental Assessment (B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711II-xxi U.S. Environmental Protection Agency Project Team

for Development of Air Quality Criteria for Lead (cont’d)

Technical Support Staff

Mr. Douglas B. Fennell—Technical Information Sp ecialist, National Center for Environmental Assessment (B243-01), U.S. Environmental Prot ection Agency, Research Triangle Park, NC 27711 (Retired)

Ms. Emily R. Lee—Management Analyst, Natio nal Center for Environmental Assessment (B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Ms. Diane H. Ray—Program Specialist, Natio nal Center for Environmental Assessment (B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Ms. Donna Wicker—Administrative Officer, Na tional Center for Environmental Assessment (B243-01), U.S. Environmental Protection Agenc y, Research Triangle Park, NC 27711 (Retired)

Mr. Richard Wilson—Clerk, National Center fo r Environmental Assessment (B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711

Document Production Staff

Ms. Carolyn T. Perry—Task Order Manager, Com puter Sciences Corpor ation, 2803 Slater Road, Suite 220, Morrisville, NC 27560

Mr. John A. Bennett—Technical Information Sp ecialist, Library Asso ciates of Maryland, 11820 Parklawn Drive, Suite 400, Rockville, MD 20852

Mr. William Ellis—Records Management Technician, InfoPro, Inc., 8200 Greensboro Drive, Suite 1450, McLean, VA 22102

Ms. Sandra L. Hughey—Technical Information Sp ecialist, Library Associates of Maryland, 11820 Parklawn Drive, Suite 400, Rockville, MD 20852

Dr. Barbara Liljequist—Technical Editor, Comp uter Sciences Corporation, 2803 Slater Road, Suite 220, Morrisville, NC 27560

Ms. Michelle Partridge-Doerr—Publications/Gra phics Specialist, TEK Systems, 1201 Edwards Mill Road, Suite 201, Raleigh, NC 27607

Mr. Carlton Witherspoon—Graphic Artist, Comput er Sciences Corporation, 2803 Slater Road, Suite 220, Morrisville, NC 27560II-xxii U.S. Environmental Protection Agency Science Advisory Board (SAB) Staff Office Clean Air Scientific Advi sory Committee (CASAC)

Chair

Dr. Rogene Henderson*—Scientist Emeritus, Lovelace Respiratory Research Institute, Albuquerque, NM

Members

Dr. Joshua Cohen—Faculty, Center for the Evaluati on of Value and Risk, Institute for Clinical Research and Health Policy Studies, Tuft s New England Medical Center, Boston, MA

Dr. Deborah Cory-Slechta—Director, University of Medicine and Dentistry of New Jersey and Rutgers State University, Piscataway, NJ

Dr. Ellis Cowling*—University Distinguished Professor-at-Large, North Carolina State University, Colleges of Natural Resources and Agriculture and Life Sciences, North Carolina State University, Raleigh, NC

Dr. James D. Crapo [M.D.]*—Professor, Departme nt of Medicine, National Jewish Medical and Research Center, Denver, CO

Dr. Bruce Fowler—Assistant Director for Scien ce, Division of Toxicology and Environmental Medicine, Office of the Director, Agency for Toxic Substances and Disease Registry, U.S. Centers for Disease Control and Pr evention (ATSDR/CDC), Chamblee, GA

Dr. Andrew Friedland—Professo r and Chair, Environmental Studies Program, Dartmouth College, Hanover, NH

Dr. Robert Goyer [M.D.]—Emeritu s Professor of Pathology, Faculty of Medicine, University of Western Ontario (Canada), Chapel Hill, NC

Mr. Sean Hays—President, Summ it Toxicology, Allenspark, CO

Dr. Bruce Lanphear [M.D.]—Sloan Professor of Children’s Environmental Health, and the Director of the Cincinnati Ch ildren’s Environmental Health Ce nter at Cincinnati Children’s Hospital Medical Center and the Univer sity of Cincinnati, Cincinnati, OH

Dr. Samuel Luoma—Senior Research Hydr ologist, U.S. Geological Survey (USGS), Menlo Park, CAII-xxiii U.S. Environmental Protection Agency Science Advisory Board (SAB) Staff Office Clean Air Scientific Advi sory Committee (CASAC) (cont’d)

Members (cont’d)

Dr. Frederick J. Miller*—Consultant, Cary, NC

Dr. Paul Mushak—Principal, PB A ssociates, and Visiting Professor, Albert Einstein College of Medicine (New York, NY), Durham, NC

Dr. Michael Newman—Professor of Marine Science, School of Marine Sciences, Virginia Institute of Marine Science, College of William & Mary, Gloucester Point, VA

Mr. Richard L. Poirot*—Environmental Analyst, Air Pollution Control Di vision, Department of Environmental Conservation, Vermont Agency of Natural Resources, Waterbury, VT

Dr. Michael Rabinowitz—Geochemist, Marine Biological Laboratory, Woods Hole, MA

Dr. Joel Schwartz—Professor, Environmental H ealth, Harvard University School of Public Health, Boston, MA

Dr. Frank Speizer [M.D.]*—Edward Kass Prof essor of Medicine, Channing Laboratory, Harvard Medical School, Boston, MA

Dr. Ian von Lindern—Senior Scientist, TerraG raphics Environmenta l Engineering, Inc., Moscow, ID

Dr. Barbara Zielinska*—Research Professor, Divi sion of Atmospheric Science, Desert Research Institute, Reno, NV

SCIENCE ADVISORY BOARD STAFF

Mr. Fred Butterfield—CASAC Designated Fede ral Officer, 1200 Pennsyl vania Avenue, N.W., Washington, DC, 20460, Phone: 202-343-9994, Fax: 202-233-0643 (butterfield.fred@epa.gov ) (Physical/Courier/FedEx Address: Fred A. Butterf ield, III, EPA Science Advisory Board Staff Office (Mail Code 1400F), Woodies Building, 1025 F Street, N.W., Room 3604, Washington, DC 20004, Telephone: 202-343-9994)

*Members of the statutory Clean Air Scientific Advisory Committee (CASAC) appointed by the U.S. EPA AdministratorII-xxiv Abbreviations and Acronyms

α FGF α -fibroblast growth factor AA arachidonic acid AAL active avoidance learning AAS atomic absorption spectroscopy ABA β -aminoisobutyric acid ACBP Achenbach Child Behavior Profile ACE angiotensin converting enzyme ACh acetylcholine AChE acetylcholinesterase ACR acute-chronic ratio AD adult ADC analog digital converter ADP adenosine diphosphate AE anion exchange AEA N -arachidonylethanolamine AFC antibody forming cells 2-AG 2-arachidonylglycerol A horizon uppermost layer of soil (litter and humus) AHR aryl hydrocarbon receptor AI angiotensin I ALA ∗ -aminolevulinic acid ALAD ∗ -aminolevulinic acid dehydratase ALAS aminolevulinic acid synthetase ALAU urinary δ -aminolevulinic acid ALD aldosterone ALS amyotrophic lateral sclerosis ALT alanine aminotransferase ALWT albumin weight AMEM Alpha Minimal Essential Medium AMP adenosine monophosphate ANCOVA analysis of covariance ANF atrial natriuretic factor Ang II angiotensin IIII-xxv ANOVA analysis of variance ANP atrial natriuretic peptide AP alkaline phosphatase AP-1 activated protein-1 ApoE apolipoprotein E AQCD Air Quality Criteria Document Arg arginine AS52 cells derived from the CHO cell line ASGP-R aceyl glycoprotein receptor AST aspartate aminotransferase ASV anode stripping voltammetry 3-AT 3-aminotriazole; 3-amino triazide ATP adenosine triphosphate ATP1A2 sodium-potassium adenosine triphosphase α 2 ATPase adenosine triphosphatase ATSDR Agency for Toxic Substances and Disease Research AVCD atrioventricular conduction deficit AVS acid volatile sulfide AWQC ambient water quality criteria ∃ beta-coefficient; slope of an equation ∃ FGF ∃ -fibroblast growth factor 17 ∃ –HS 17 ∃ -hydroxysteriod 3 ∃ -HSD 3 ∃ -hydroxysteriod dehydrogenase 17 ∃ -HSDH 17 ∃ -hydroxysteriod dehydrogenase 6 ∃ -OH-cortisol 6- ∃ -hydroxycortisol B both BAEP brainstem auditory-evoked potentials BAER brainstem audito ry-evoked responses BAF bioaccumulation factor B cell B lymphocyte BCFs bioconcentration factors BCS bovine calf serum BDNF brain derived neurotrophic factor BDWT body weight changes BEI biological exposure indexII-xxvi BFU-E blood erythroid progenitor BLL blood lead level BLM biotic ligand model BM basement membrane BMI body mass index BDNF brain-derived ne urotrophic factor BOTMP Bruinicks-Oseretsky Test of Motor Proficiency BP blood pressure BPb blood lead concentration BSA bovine serum albumin BSI Brief Symptom Inventory BTQ Boston Teacher Questionnaire BUN blood urea nitrogen bw, b. wt., BW body weight C3H10T/12 mouse embryo cell line C3, C4 complement proteins CA chromosome aberration CA3 cornu ammonis 3 region of hippocampus

45 Ca calcium-45 radionuclide Ca-ATP calcium-dependent adenosine triphosphate Ca-ATPase calcium-dependent adenosine triphosphatase CaCO 3 calcium carbonate CaEDTA calcium disodium ethyl enediaminetetraacetic acid CAL calcitonin CaM calmodulin Ca-Mg-ATPase calcium-magnesium-dependent adenosine triphosphatase cAMP cyclic adenosinemonophosphate CaNa 2 EDTA calcium disodium ethyl enediaminetetraacetic acid CANTAB Cambridge Neuropsychologi cal Testing Automated Battery CAT catalase; Cognitive Abilities Test CBCL Achenbach Child Behavior Checklist CBCL-T Total Behavior Problem Score CBL cumulative blood lead CBLI cumulative blood lead index CCB cytochalasin BII-xxvii CCD charge-coupled device CCE Coordination Center for Effects CCL carbon tetrachloride CCS cosmic calf serum C-CV RSA coefficient of component variance of respiratory sinus arrhythmia Cd cadmium

109 Cd cadmium-109 radionuclide CdU urinary cadmium CEC cation exchange capacity CESD, CES-D Center for Epidemio logic Studies Depression (scale) GFAP glial fibrillary acidic protein CFU-E colony forming unit blood-er ythroid progenitor (cell count) CFU-GEMM colony forming unit blood- pluripotent progen itor (cell count) CFU-GM blood granulocyte/macro phage progenitor (cell count) cGMP cyclic guanosine-3',5'-monophosphate ChAT choline acetyltransferase CHD coronary heart disease CHO Chinese hamster ovary cell line CI confidence interval CLE-SV competitive ligand-exchange/stripping voltammetry CLRTAP Convention on Long-Range Transboundary of Air Pollution CLS Cincinnati Lead Study CMC criterion maximum concentration CMI cell-mediated immunity CNS central nervous system COH cation-osmotic hemolysis ConA concanavalin A COR cortisol CoTx cotreatment COX-2 cyclooxygenase-2 CP coproporphryn CPT current perception threshold cr creatinine CRAC calcium release activated calcium reflux CREB cyclic AMP-response element binding proteinII-xxviii CRF chronic renal failure CRI chronic renal insufficiency CSF cerebrospinal fluid CuZn-SOD copper and zinc-dependent superoxide dismutase CV conduction velocity CVLT California Verbal Learning Test CV R-R coefficient of variation of the R-R interval CYP cytochrome (e.g., CYP1A, CYP-2A6, CYP3A4, CYP450) CYP3a11 cytochrome P450 3a11 D D-statistic DA dopamine; dopaminergic dbcAMP dibutyryl cyclic ad enosine-3',5'-monophosphate DCV distribution of conduction velocities DEAE diethylaminoet hyl (chromatography) DET diffusive equilibrium thin films DEYO death of young DFS decayed or filled surfaces, permanent teeth dfs covariate-adjusted number of caries DG dentate gyrus DGT diffusive gradient thin films DL DL-statistic DMEM Dulbecco’s Minimal Essential Medium DMEM/F12 Dulbecco’s Minimal Essential Medium/Ham’s F12 DMFS decayed, missing, or filled surfaces, permanent teeth DMPS 2,3-dimercaptopropane 1-sulfonate DMSA 2,3-dimercaptosuccinic acid DMT Donnan membrane technique DMTU dimethylthiourea DNA deoxyribonucleic acid DO distraction osteogenesis DOC dissolved organic carbon DOM dissolved organic carbon DOPAc 3,4-dihydroxyphenylacetic acid DPASV differential pulse anod ic stripping voltammetry dp/dt rate of left ventri cular isovolumetric pressureII-xxix DPPD N

- N -diphenyl- p -phynylene-diamine DR drinking water DSA delayed spatial alternation DTC diethyl dithiocarbomate complex DTH delayed type hypersensitivity DTPA diethylenetriaminepentaacetic acid DTT dithiothreitol dw dry weight E embryonic day E 2 estradiol EBE early biological effect EBV Epstein-Barr virus EC European Community EC 50 effect concentration fo r 50% of test population eCB endocannabinoid ECG electrocardiogram Eco-SSL ecological soil screening level EDS energy dispersive spectrometers EDTA ethylenediaminetetraacetic acid EEDQ N -ethoxycarbonyl-2-eth oxy-1,2-dihydroquinone EEG electroencephalogram EG egg EGF epidermal growth factor EGG effects on eggs EGPN egg production EKG electrocardiogram electro electrophysiological stimulation EM/CM experimental medium-to-control medium (ratio) EMEM Eagle’s Minimal Essential Medium eNOS endothelial nitr ic oxide synthase EP erythrocyte protoporphyrin EPA U.S. Environmental Protection Agency Epi epinephrine EPMA electron probe microanalysis EPO erythropoietinII-xxx EPSC excitatory postsynaptic currents EPT macroinvertebrates from the Ephemeroptera (mayflies), Plecoptera (stoneflies), and Trichoptera (caddisflies) group ERG electroretinogram; electroretinographic ERL effects range – low ERM effects range – median EROD ethoxyresorufin- O -deethylase ESCA electron spectroscopy for chemical analysis ESRD end-stage renal disease EST estradiol ESTH eggshell thinning ET endothelein; essential tremor ETOH ethyl alcohol EXAFS extended X-ray absorption fine structure EXANES extended X-ray absorption near edge spectroscopy F F-statistic F344 Fischer 344 (rat) FAV final acute value FBS fetal bovine serum FCS fetal calf serum FCV final chronic value FD food FEF forced expiratory flow FEP free erythrocyte protoporphyrin FERT fertility FEV 1 forced expiratory volume in one second FGF fibroblast growth factor (e.g., β FGF, α FGF) FI fixed interval (operant conditioning) FIAM free ion activity model FMLP N -formyl-L-methionyl-L-l eucyl-L-phenylalanine fMRI functional magnetic resonance imaging FR fixed-ratio operant conditioning FSH follicle stimulating hormone FT3 free triiodothyronine FT4 free thyroxineII-xxxi FTES free testosterone FTII Fagan Test of Infant Intelligence FTPLM flow-through permeation liquid membranes FURA-2 1-[6-amino-2-(5-carboxy-2- oxazolyl)-5-benzofuranyloxy]-2-(2- amino-5-methylphenoxy) ethane- N , N , N' , N' -tetraacetic acid FVC forced vital capacity ( -GT ( -glutamyl transferase G gestational day GABA gamma aminobutyric acid GAG glycosaminoglycan G12 CHV79 cells derived from the V79 cell line GCI General Cognitive Index GD gestational day GDP guanosine diphosphate GEE generalized estimating equations GFAAS graphite furnace atomic absorption spectroscopy GFR glomerular filtration rate GGT ( -glutamyl transferase GH growth hormone GI gastrointestinal GIME-VIP gel integrated microelect rodes combined with voltammetric in situ profiling GIS geographic information system GLU glutamate GMAV genus mean acute value GMCV genus mean chronic value GMP guanosine monophosphate GMPH general morphology GnRH gonadotropin releasing hormone GOT aspartate aminotransferase GP gross productivity G6PD, G6PDH glucose-6- phosphate dehydrogenase GPEI glutathione S -transferase P enhancer element gp91 phox NAD(P)H oxidase GPT glutamic-pyruvic transaminase GPx glutathione peroxidaseII-xxxii GRO growth GRP78 glucose-regulated protein 78 GSD geometric standard deviation GSH reduced glutathione GSIM gill surface interaction model GSSG glutathione disulfide GST glutathione- S -transferase GSTP placental glutathione transferase GTP guanosine triphosphate GV gavage H + acidity

3 H hydrogen-3 radionuclide (tritium) HA humic acid; hydroxyapatite Hb hemoglobin HBEF Hubbard Brook Experimenatl Forest HBSS Hank’s Balanced Salt Solution HCG; hCG human chorionic gonadotropin Hct hematocrit HDL high-density lipopr otein (cholesterol) HEP habitat evaluation procedure HET Binghamton heterogeneous stock HFPLM hollow fiber permeation liquid membranes Hgb hemoglobin HGF hepatocyte growth factor HH hydroxylamine hydrochloride H-H high-high HHANES Hispanic Health and Nutrition Examination Survey H-L high-low HLA human leukocyte antigen H-MEM minimum essential medium/nutrient mixture–F12-Ham HMP hexose monophosphate shunt pathway HNO 3 nitric acid H 2 O 2 hydrogen peroxide HOME Home Observation for Measurement of Environment HOS TE human osteosarcoma cellsII-xxxiii HPLC high-pressure liquid chromatography H 3 PO 4 phosphoric acid HPRT hypoxanthine phosphoribosyltransferase (gene) HR heart rate HSI habitat suitability indices H 2 SO 4 sulfuric acid HSPG heparan sulfate proteoglycan Ht hematocrit HTC hepatoma cells hTERT catalytic subunit of human telomerase HTN hypertension IBL integrated blood lead index

IBL Η WRAT-R integrated blood lead index Η Wide Range Achievement Test-Revised (interaction) ICD International Classification of Diseases ICP inductively coupled plasma ICP-AES inductively coupled plas ma atomic emission spectroscopy ICP-MS, ICPMS inductively coupl ed plasma mass spectrometry ID-MS isotope dilution mass spectrometry IFN interferon (e.g., IFN- ( ) Ig immunoglobulin (e.g., IgA, IgE, IgG, IgM) IGF-1 insulin-like growth factor 1 IL interleukin (e.g., IL-1, IL-1 ∃ , IL-4, IL-6, IL-12) ILL incipient lethal level immuno immunohistochemical staining IMP inosine monophosphate iNOS inducible nitric oxide synthase i.p., IP intraperitoneal IPSC inhibitory postsynaptic currents IQ intelligence quotient IRT interresponse time ISEL in situ end labeling ISI interstimulus interval i.v., IV intravenous IVCD intraventricular conduction deficitII-xxxiv JV juvenile KABC Kaufman Assessment Battery for Children KTEA Kaufman Test of Educational Achievement KXRF, K-XRF K-shell X-ray fluorescence LA lipoic acid LB laying bird LC lactation LC 50 lethal concentration at which 50% of exposed animals die LC 74 lethal concentration at which 74% of exposed animals die LD 50 lethal dose at which 50% of exposed animals die LDH lactate dehydrogenase LDL low-density lipoprotein (cholesterol) L-dopa 3,4-dihydroxyphenylalanine (precursor of dopamine) LE Long Evans (rat) LET linear energy transfer (radiation) LH luteinizing hormone LHRH luteinizing hormone releasing hormone LN lead nitrate L-NAME L- N G -nitroarginine methyl ester LOAEL lowest-observed adverse effect level LOEC lowest-observed-effect concentration LOWESS locally weighted scatter plot smoother LPO lipoperoxide LPP lipid peroxidation potential LPS lipopolysaccharide LT leukotriene LT 50 time to kill 50% LTER Long-Term Ecological Research (sites) LTP long term potentiation LVH left ventricular hypertrophy μPIXE microfocused particle induced X-ray emission μSXRF microfocused synchrotron-based X-ray fluorescence MA mature MA-10 mouse Leydig tumor cell line MANCOVA multivariate analysis of covarianceII-xxxv MAO monoamine oxidase MATC maximum acceptable threshold concentration MDA malondialdehyde MDA-TBA malondialdehyde-thiobarbituric acid MDCK kidney epithelial cell line MDI Mental Development Index (score) MDRD Modification of Diet in Renal Disease (study) MEM Minimal Essential Medium MG microglobulin Mg-ATPase magnesium-dependent adenosine triphosphatase MiADMSA monoisamyl dimercaptosuccinic acid Mi-DMSA mi monoisoamyl dimercaptosuccinic acid MK-801 NMDA receptor antagonist MLR mixed lymphocyte response MMSE Mini-Mental State Examination MMTV murine mammary tumor virus MN micronuclei formation MND motor neuron disease MNNG N -methyl- N' -nitro- N -nitrosoguanidine MPH morphology MRI magnetic resonance imaging mRNA messenger ribonucleic acid MROD methoxyresorufin- O -demethylase MRS magnetic resonance spectroscopy MS mass spectrometry MSCA McCarthy Scales of Children’s Abiltities mSQGQs mean sediment quality guideline quotients MT metallothionein MVV maximum voluntary ventilation MW molecular weight (e.g., high-MW, low-MW) N, n number of observations N/A not available NAAQS National Ambient Air Quality Standards NAC N -acetyl cysteine NAD nicotinamide adenine dinucleotideII-xxxvi NADH reduced nicotinamide adenine dinucleotide NADP nicotinamide adenine dinucleotide phosphate NAD(P)H, NADPH reduced nicotinamide adenine dinucleotide phosphate NADS nicotinamide adenine dinucleotide synthase NAF nafenopin NAG N -acetyl- ∃ -D-glucosaminidase Na-K-ATPase sodium-potassium-dep endent adenosine triphosphatase NAWQA National Water-Quality Assessment NBT nitro blue tetrazolium NCBP National Contaminant Biomonitoring Program NCD nuclear chromatin decondensation (rate) NCS newborn calf serum NCTB Neurobehavioral Core Test Battery NCV nerve conduction velocity ND non-detectable; not detected NDI nuclear divison index NE norepinephrine NES Neurobehavioral Evaluation System NF- κ B nuclear transcription factor- κ B NGF nerve growth factor NHANES National Health and Nutrition Examination Survey NIOSH National Institute for Occupational Safety and Health NIST National Institute for Standards and Technology NK natural killer NMDA N -methyl-D-aspartate NMDAR N -methyl-D-aspartate receptor NMR nuclear magnetic resonance NO nitric oxide NO 2 nitrogen dioxide NO 3 nitrate NOAEC no-observed-adverse- effect concentration NOAEL no-observed-adverse-effect level NOEC no-observed-effect concentration NOEL no-observed-effect level NOM natural organic matterII-xxxvii NORs nucleolar organizing regions NOS nitric oxide synthase ; not otherwise specified NO x nitrogen oxides NP net productivity NPSH nonprotein sulfhydryl NR not reported NRC National Research Council NRK normal rat kidney NS nonsignificant NSAID non-steroidal anti-inflammatory agent NT neurotrophin NTA nitrilotriacetic acid O 2 oxygen ODVP offspring development OH hydroxyl 7-OH-coumarin 7-hydroxy-coumarin 1,25-OH-D, 1,25-OH D 3 1,25-dihydroxyvitamin D 24,25-OH-D 3 24,25-dihydroxyvitamin D 25-OH-D 3 25-hydroxyvitamin D 8-OHdG 8-hydroxy-2'-deoxyguanosine O horizon forest floor OR odds ratio; other oral OSWER Office of Solid Waste and Emergency Response P, p probability value P300 event-related potential P450 1A1 cytochrome P450 1A1 P450 1A2 cytochrome P450 1A2 P450 CYP3a11 cytochrome P450 3a11 PAD peripheral arterial disease PAH polycyclic aromatic hydrocarbon PAI-1 plasminogen activator inhibitor-1 PAR population attributable risk Pb lead

203 Pb lead-203 radionuclide

204 Pb, 206 Pb, 207 Pb, 208 Pb stable isotopes of lead -204, -206, -207, -208, respectivelyII-xxxviii 210 Pb lead-210 radionuclide Pb(Ac) 2 lead acetate PbB blood lead concentration PbCl 2 lead chloride Pb(ClO 4 ) 2 lead chlorate PBG-S porphobilinogen synthase PBMC peripheral bl ood mononuclear cells Pb(NO 3 ) 2 lead nitrate PbO lead oxides (or litharge) PBP progressive bulbar paresis PbS galena PbU urinary lead PC12 pheochromocytoma cell PCR polymerase chain reaction PCV packed cell volume PDE phosphodiesterase PDGF platelet-derived growth factor PDI Psychomotor Development Index PEC probable effect concentration PEF expiratory peak flow PG prostaglandin (e.g., PGE 2 , PGF 2 ); prostate gland PHA phytohemagglutinin A Pi inorganic phosphate PIXE particle induced X-ray emission PKC protein kinase C pl NEpi plasma norepinephrine PMA progressive muscular atrophy PMN polymorphonuclear leucocyte PMR proportionate mortality ratio PN postnatal (day) P5N pyrimidine 5'-nucleotidase PND postnatal day p.o., PO per os (oral administration) POMS Profile of Mood States ppb parts per billionII-xxxix ppm parts per million PPVT-R Peabody Picture Vocabulary Test-Revised PRA plasma renin activity PRL prolactin PROG progeny counts or numbers PRR prevalence rate ratio PRWT progeny weight PST percent transferrin saturation PTH parathyroid hormone PTHrP parathyroid hor mone-related protein PVC polyvinyl chloride PWM pokeweed mitogen PRWT progeny weight QA/QC quality assura nce/quality control Q/V flux of air (Q) divided by volume of culture (V) r Pearson correlation coefficient R 2 multiple correlation coefficient

r 2 correlation coefficient

226 Ra most stable isotope of radium R/ALAD ratio of ALAD activity be fore and after reactivation RAVLT Rey Auditory Verbal Learning Test

86 Rb rubidium-86 radionuclide RBA relative bioavailablity RBC red blood cell; erythrocyte RBF renal blood flow RBP retinol binding protein RBPH reproductive behavior RCPM Ravens Colored Progressive Matrices REL rat epithelial (cells) REP reproduction RHIS reproductive organ histology

222 Rn most stable isotope of radon RNA ribonucleic acid ROS reactive oxygen species ROS 17.2.8 rat osteosarcoma cell lineII-xl RPMI 1640 Roswell Park Memorial In stitute basic cell culture medium RR relative risk; rate ratio RT reaction time RSEM resorbed embryos RSUC reproductive success (general) RT reproductive tissue ∑ SEM sum of the molar concentrations of simultaneously extracted metal SA7 simian adenovirus SAB Science Advisory Board SAM S -adenosyl-L-methionine SBIS-4 Stanford-Binet Intelligence Scale-4th edition s.c., SC subcutaneous SCAN Test for Auditory Processing Disorders SCE selective chemical extraction; sister chromatid exchange SCP stripping chronopotentiometry SD Spraque-Dawley (rat); standard deviation SDH succinic acid dehydrogenase SDS sodium dodecyl sulfate; Symbol Digit Substitution SE standard error; standard estimation SEM standard error of the mean SES socioeconomic status sGC soluble guanylate cyclase SH sulfhydryl SHBG sex hormone binding globulin SHE Syrian hamster embryo cell line SIMS secondary ion mass spectrometry SIR standardized incidence ratio SLP synthetic leaching procedure SM sexually mature SMAV species mean acute value SMR standardized mortality ratio SNAP Schneider Neonatal Assessment for Primates SNP sodium nitroprusside SO 2 sulfur dioxide SOD superoxide dismutaseII-xli SOPR sperm-oocyte penetration rate SPCL sperm cell counts SPCV sperm cell viability SQGs sediment quality guidelines SRA Self Reported Antisocial Behavior scale SRD Self Report of Delinquent Behavior SRIF somatostatin SRM Standard Reference Material SRT simple reaction time SSADMF Social Security Admini stration Death Master File SSB single-strand breaks SSEP somatosensory-evoked potential StAR steroidogenic acute regulatory protein STORET STOrage and RETrieval SVC sensory conduction velocity SVRT simple visual reaction time T testosterone TA tail TABL time-averaged blood lead T&E threatened and endangered (species) TAT tyrosine aminotransferase TB tibia TBARS thiobarbituric acid-reactive species TBPS Total Behavior Problem Score TCDD methionine-choline-deficient diet T cell T lymphocyte TCLP toxic characteristic leaching procedure TE testes TEC threshold effect concentration TEDG testes degeneration TEL tetraethyl lead TES testosterone TEWT testes weight TF transferrin, translocation factor TG 6-thioguanineII-xlii TGF transforming growth factor TH tyrosine hydroxylase

232 Th stable isotope of thorium-232 TLC Treatment of Lead-exposed Children (study) TNF tumor necrosis factor (e.g., TNF- α ) TOF time-of-flight tPA plasminogen activator TPRD total production TRH thyroid releasing hormone TRV toxicity reference value TSH thyroid stimulating hormone TSP triple-super phosphate TT3 total triiodothyronine TT4 serum total thyroxine TTES total testosterone TTR transthyretin TU toxic unit TWA time-weighted average TX tromboxane (e.g., TXB 2 ) U uriniary

235 U, 238 U uranium-234 and -238 radionuclides UCP urinary coproporphyrin UDP uridine diphosphate UNECE United Nations Economic Commission for Europe Ur urinary USFWS U.S. Fish and Wildlife Service USGS United States Geological Survey UV ultraviolet V79 Chinese hamster lung cell line VA Veterans Administration VC vital capacity; vitamin C VDR vitamin D receptor VE vitamin E VEP visual-evoked potential VI variable-intervalII-xliii vit C vitamin C vit E vitamin E VMA vanilmandelic acid VMI Visual-Motor Integration VSM vascular smooth muscle (cells) VSMC vascular smooth muscle cells WAIS Wechsler Adult Intelligence Scale WDS wavelength dispersive spectrometers WHO World Health Organization WISC Wechsler Intelligence Scale for Children WISC-R Wechsler Intelligence Scale for Children-Revised WO whole organism WRAT-R Wide Range Achievement Test-Revised WT wild type WTHBF-6 human liver cell line ww wet weight XAFS X-ray absorption fine structure XANES X-ray absorption near edge spectroscopy XAS X-ray absorption spectroscopy XPS X-ray photoelectron spectroscopy X-rays synchrotron radiation XRD X-ray diffraction XRF X-ray fluorescence ZAF correction in reference to thr ee components of matrix effects: atomic number (Z), absorpti on (A), and fluorescence (F) ZnNa 2 DTPA zinc disodium diet hylenetriaminepentaacetic acid ZnNa 2 EDTA zinc disodium ethyl enediaminetetraacetic acid ZPP zinc protoporphyrinAX7-1 

## AX7. CHAPTER 7 ANNEX - ENVIRONMENTAL EFFECTS OF LEAD

AX7.1 TERRESTRIAL ECOSYSTEMS AX7.1.1 Methodologies Used in Terrestrial Ecosystems Research The distribution of Pb throughout the terrestrial ecosystem, via aerial deposition, has been discussed throughout this document. Its furthe r impacts on soil, sediment, and water provide numerous pathways that may promote unacceptable risk to all levels of biota. Stable isotopes of Pb have been found useful in identifying sources and apportionment to various sources. One of the key factors affecting assessment of risk is an understanding, and pe rhaps quantification, of bioavailability. Therefore, the bioavailability of Pb is a key issue to the development of NAAQS. However, the discussion of all methods used in characterizi ng bioavailability is beyond the scope of this chapter. The following topics are discussed in this chapter.

• Lead Isotopes and Apportionment

• Methodologies to determine Pb speciation

• Lead and the Biotic Ligand Model (BLM)

• In situ methods to reduce Pb bioavailability

AX7.1.1.1 Lead Isotopes and Apportionment

Determination of the extent of Pb contamin ation from an individual source(s) and its impact are of primary importance in risk assessm ent. The identification of exposure pathway(s)

is fundamental to the risk anal ysis and critical in the planni ng of remediation scenarios. Although societies have been consuming Pb fo r nearly 9,000 years, production of Pb in the United States peaked in 1910 and 1972, at ap proximately 750 and 620 kt/year, respectively (Rabinowitz, 2005). The diversity of potential Pb sources (fossil fuel burning, paint pigments, gasoline additives, solders, ceramics, batteries ) and associated production facilities (mining, milling, smelting-refining) make fingerprinting of sources difficult. (See Chapter 2 and its Annex for additional information on sources.) Therefore, dealing with multiple sources (point and nonpoint), a reliable and speci fic fingerprinting technique is required. It has been well established (Sturges and Barrie, 1987; Rabinowitz, 1995) that the stable isotope composition ofAX7-2 Pb is ideally suited for this task. Lead isotopic ratio differences often allow multiple sources to be distinguished, with an apportionment of the bu lk Pb concentration made to those sources. Lead has four stable isotopes: 204 Pb, 206 Pb, 207 Pb, and 208 Pb in natural abundances of 1.4, 24.1, 22.1, and 52.4%, respectively. The radiogenic 206 Pb, 207 Pb, and 208 Pb are produced by radioactive decay of 238 U, 235 U, and 232 Th, respectively. Thus, the isotopic composition of Pb

varies based on the U:Pb and Th:Pb ratios of th e original ore’s source and age (Faure, 1977).

Because of the small fractional mass differences of the Pb isotopes, ordinary chemical and pyrometallurgical reactions will not alter their original composition. Therefore, anthropogenic sources reflect the isotopic composition of the ores from which the Pb originated. To acquire the Pb isotopes, a sample, gene rally in aqueous form, is analyzed on an ICP/MS (quadrapole, magnetic sector, or time-o f-flight). Studies reviewing the most common analytical and sample preparati on procedures include Ghazi and Millette (2004), Townsend et al. (1998), and Encinar et al. (2001a,b) . The correction factor for mass discrimination biases is generally made by analyzing the National In stitute for Standards and Technology (NIST),

Standard Reference Material (SRM) 981 and/or spiked 203 Tl and 205 Tl isotopes (Ketterer et al., 1991; Begley and Sharp, 1997). The overall succe ss of Pb isotope fingerprinting is generally dependent on analysis precision, which in turn depends on the type of mass analyzer used (Table AX7-1.1.1).

Table AX7-1.1.1. Relative Standard Deviation (RSD) for Lead Isotope Ratios on Selected Mass Spectrometers

RSD Quadrapole Double-Focusing Single-Focusing Magnetic Sector High-Resolution Magnetic Sector ICP/MS

204 Pb: 206 Pb 0.0031 0.0032 0.00053 0.0011

207 Pb: 206 Pb 0.0032 0.0027 0.00053 0.00048

208 Pb: 206 Pb 0.0026 0.0024 0.00053 0.00046

An extensive database comprising primarily No rth American Pb sources can be assembled from Doe and Rohrbough (1977), Doe and Stacey (1974) , Doe et al. (1968), Heyl et al. (1974),AX7-3 Leach et al. (1998), Stacey et al. (1968), Zartman (1974), Cannon and Pierce (1963), Graney et al. (1996), Unruh et al . (2000), James and Henry ( 1993), Rabinowitz (2005), and Small (1973). The use of Pb isotopes to quantitatively appor tion source contributions follows the simple mixing rule when only two sources are possible (Faure, 1977). Once multiple sources need to be considered, a unique solution can no longer be calculated (Fry and Sherr, 1984). Phillips and Gregg (2003) have designed a model to give feasible source contributions when multiple sources are likely. Many studies have demonstrated the usefulness of this apportionment technique. Media of all types have been st udied: water (Flegal et al., 1989a,b; Erel et al., 1991; Monna et al., 1995), ice (Planchon et al., 2002), dust (Adgate et al., 1998; Sturges et al., 1993), and soil/sediments (Hamelin et al., 1990; Farmer et al ., 1996; Bindler et al., 1999; Haack et al., 2004; Rabinowitz and Wetherill, 1972; Rabi nowitz, 2005; Kettere r et al., 2001).

AX7.1.1.2 Speciation in Assessing Lead Bioava ilability in the Terres trial Environment One of the three processes defined by the National Research Council in its review on bioavailability (NRC, 2002) is “contaminant in teractions between phases”, more commonly referred to as “speciation.” A wide variety of analytical (XRD, EP MA, EXAFS, PIXE, XPS, XAS, SIMS) and chemical speciation modeling (SOILCHEM, MINTEQL, REDEQL2, ECOSAT, MINTEQA2, HYDRAQL, PHREEQE, WATEQ4F) t ools have been used to char acterize a metal’s speciation as it is found in various media. Currently, for risk assessment purposes (not considering phytotoxicity), where large sites with numerous media, pathways, and metals must often be characterized in a reasonable time frame, electron microprobe analysis (EMPA) techniques provide the greatest information on metal speciati on. Other techniques such as extended X-ray absorption fine structure (EXAFS) and exte nded X-ray absorption near edge spectroscopy (EXANES) show great promise and will be import ant in solving key mechanistic questions. In the case of phytotoxicity, the speciation of me tals by direct measurement or chemical models of pore-water chemistry is most valuable. Fu rther work needs to be done in developing analytical tools for the speciation of the methyl -forming metals (Hg, As, Sb, Se, and Sn) in soils

and sediments.AX7-4 Concept For a given metal or metalloid (hereafter referred to as metal), the term speciation refers to its chemical form or species, including its phys icochemical characteristics that are relevant to bioavailability. As a result of the direct im pact these factors often have on a metal’s bioavailability, the term “bioaccessibility” ha s been adopted to define those factors.

Speciation Role The accumulation of metals in the lithosphe re is of great concern. Unlike organic compounds, metals do not degrade and, thus, have a greater tendency to bioaccumulate. It is now well accepted that knowledge of the bulk, toxic characteris tic leaching procedure (TCLP), or synthetic leaching procedure (SLP) concentrations for any metal is not a controlling factor in understanding a metal’s environmental behavior or in developing remedies for its safe management. Although these tests are essential to site characterization and management, they offer no insight into risk assessment. Rather, it is the metal’s bioavailabil ity (the proportion of a toxin that passes a physiological membrane [the pl asma membrane in plants or the gut wall in animals] and reaches a target receptor [cytosol or blood]), which plays a significant role in the

risk assessment of contaminated media. The NRC review (NRC, 2002) on bioavailability defined bioavailability processes in terms of three key processes:

• contaminant interactions be tween phases (association- dissociation/bound-released),

• transport of contaminants to organism, and

• passage across a physiological membrane.

As mentioned previously, the first process is more commonl y referred to as speciation. The speciation of a toxic metal in the environm ent is a critical com ponent of any ecosystem health risk assessment. Four important toxi cologic and toxicokinetic determinants relating speciation to bioavailability are the (1) chemical form or species, (2) part icle size of the metal form, (3) lability of the ch emical form, and (4) source.AX7-5 Chemical Form of Species The solid phase in a medium controls the activity of a metal in solution, whether the solution is surface, ground, or pore water or GI fluids, and plays a profound role in metal bioavailability. This is perhaps best illustrate d by in vivo and in vitro results for many of the common Pb-bearing minerals (D rexler, 1997) (Figure AX7-1.1.1). The metal species found in media are often diverse, and data suggest that their bioavailability may be significantly influenced by site-specific variations within th ese identified metal speci es (Davis et al., 1993; Ruby et al., 1992; Drexler and Mushak, 1995).

Figure AX7-1.1.1. Relationship of bioaccessibili ty (low, medium, high) versus speciation as shown with scanning electron micrographs of various Pb-bearing materials.

Particle Size of Metal Species Particle size of a metal form is an important factor in the mobilization of the metal, primarily because as size decreases, the surface area of the particle increases, thereby increasing solubility. Thus, although solubi lity is not the only control fo r bioavailability, an increase in bioavailability has been directly attributed to a decrease in pa rticle size: Barltrop and MeekAX7-6 (1979) observed that “the smaller the lead pa rticle, the higher blood lead level.” Similar observations were made by Healy et al. (1992) using and an in vitro dissolution technique. Drexler (1997) presented in vi tro results on numerous Pb-bearing phases ranging in particle size from 35 to 250 μm. While all phases studied show ed increased bioavailab ility with decreasing particle size, more significantly, not all forms showed the same de gree or magnitude of change (Figure AX7-1.1.2). Atmospheric particles are ge nerally found to occur in bimodal populations: fine; 0.1 to 2.5 μm and coarse; 2.5 to 15 μm. This distribution is bot h a function of there transport mechanism and emission source. Although the upper size limit for particles that can be suspended in air is about 75 μm (Cowherd et al ., 1974), other means of mechanical entrainment (saltation, and creep) can transpor t particles as large as 1000 μm , supporting the importance of fugative emissions on media contamination. In addition, particle size can change post depositional, as soluble forms re-precipitate or sorb onto other surfaces. Limited data are available on the particle-size of discrete Pb phases from multimedia environments. One example is the study by Drexler, 2004 at Herculaneum, Miss ouri. At this site, galena (PbS) was the dominant Pb species with mean particle-s ize distributions of 4, 6, and 14 μm in PM 10 filters, house dust, and soils, respectively. These findings support the conclusion that aerial transport was the primary mechanism for Pb deposition in resi dential yards. Finally, such laboratory data have been supported by extensive epidemiologic evidence, enforci ng the importance of particle size (Bornschein et al., 1 987; Brunekreef et al., 1983; Angle et al., 1984).

Particle Lability The impact on bioavailability of a metal partic le’s lability (its associations within the medium matrix) is not well documented, but it follows the premise put forth by many of the developing treatment technologies regarding its being bound or isol ated from its environment. Data from several EPA Superfund sites and th e Region VIII swine study (U.S. Environmental Protection Agency, 2004a) suggest that matrix asso ciations, such as liberated versus enclosed, can play an important part in bioavailability. As illustrate d in Figure AX7-1.1.3, two different

media with similar total Pb concentrations a nd Pb forms (slag, Pb-oxide, and Pb-arsenate) exhibit significantly different bioavailabilities. In the Murray, UT sample (bioaccumulation factor [BAF] = 53%), a greater fraction of the more bioavailable Pb-oxides are liberated and not enclosed in the less-soluble glass-like slag as observed in the Leadville, CO sampleAX7-7 Figure AX7-1.1.2. Variation of bioavailability with particle size.

(BAF = 17%). Other evidence is more empiri cal, as illustrated in Figure AX7-1.1.4, where a large particle of native Pb is shown to have developed a weathering ring of highly bioavailable

Pb-chloride and Pb-oxide. Such observations ca n be useful in understanding the mechanistic phenomena controlling bioavailability. In additi on, they will aid in developing and validating models to predict metal-e nvironment interactions.AX7-8 Figure AX7-1.1.3. Illustration of particle lability and bioavailab ility at two different sites with similar total Pb concentrations and Pb forms.

Figure AX7-1.1.4. Scanning electron micrograph of a large native Pb particle showing outer ring of highly bioavaila ble Pb-chloride and Pb-oxide.AX7-9 Source Although the source of a metal is not directly related to bioavailability, it plays an important role in risk assessment with the evaluation of metal (1) pathways, (2) background, and (3) apportionment. It is important to understand a metal’s pathway before any remedial action can be taken; otherwise, recontamina tion of the primary pathway and reexposure can occur. Knowledge of background is important, as an action level cannot be established below natural background levels.

Plants When considering the bioavailabil ity of a metal to plants from soils and sediments, it is generally assumed that both the kinetic rate of su pply and the speciation of the metal to either the root or shoot are highly important. In soils and sediments generally, only a small volume of

water is in contact with the chemical form, and although the proportion of a metal’s concentration in this pore water to the bulk soil/s ediment concentration is small, it is this phase that is directly availabl e to plants. Therefore, pore water chemistry (i.e., metal concentration as simple inorganic species, organic complexes, or colloid complexes) is most important. Tools currently used for metal speciation fo r plants include (1) in-situ measurements using selective electrodes (Gundersen et al., 199 2; Archer et al., 1989; Wehrli et al., 1994); (2) in-situ collection techniques using diffusive equilibrium thin films (DET) and diffusive gradient thin films (DGT) followed by laborator y analyses (Davison et al., 1991, 1994; Davison and Zhang, 1994; Zhang et al., 1995); and (3) equilibrium models ( SOILCHEM) (Sposito and Coves, 1988).

AX7.1.1.3 Tools for Bulk Lead Qu antification and Speciation Bulk Quantification The major analytical methods most commonly us ed for bulk analyses outlined in the 1986 Lead ACQD included:

• Atomic Absorption Spectrometry (AAS)

• Emission Spectrometry (Inductively coupled plasma/atomic emission spectrometry)

• X-ray Fluorescence (XRF)

• Isotope Dilution Mass Spectrometry (ID/MS)AX7-10 • Colorimetric

• Electrochemical (anodic stri pping voltametry and differe ntial pulse polarography).

The choice of analytical method today for bul k quantification is generally ICP/AES or ICP/MS (U.S. Environmental Protection Ag ency, 2001). Since 1986, numerous SRMs have been developed for Pb (Table AX7-1.1.2), and se veral significant tec hnological improvements have been developed.

Table AX7-1.1.2. National Institute of Standards and Technology Lead SRMs

NIST SRM Medium Mean Pb mg/kg 2710 Soil 5532

2711 Soil 1162

2709 Soil 18.9 2587 Soil (paint) 3242 2586 Soil (paint) 432 2783 Filter (PM 2.5 ) 317 1648 Urban particulate 6550

1649a Urban dust 12,400

2584 Indoor dust 9761

2583 Indoor dust 85.9 1515 Apple leaves 0.47

1575 Pine needles 0.167

Modern spectrometry systems have replaced photomultiplier tubes with a charge-coupled device (CCD). The CCD is a camera that can de tect the entire light spectrum (>70,000 lines) from 160 to 785 nm. This allows for the simultaneous measurement of all elements, as well as any interfering lines (a productiv ity increase), and increases precis ion. The detection limit for Pb in clean samples can now be as low as 40 ppb.AX7-11 Modern ICP/AES systems offer a choice of e ither axial viewed pl asma (horizontal), which provides greater sensitivity (DL= 0.8 μg Pb/L ), or radial (vertical) viewed plasma, which performs best with high total disso lved samples (DL = 5.0 μg Pb/L). The development of reaction or collision cells have expanded the capabilities of ICP/MS and lowered detection limits for many elements that were difficult to analyze because of interferences such as Se, As, Ti, Zn, Ca, Fe, a nd Cr. The cells provide efficient interference (isobaric, polyatomic, and argi de) removal independent of the analyte and sample matrix by using various reaction gases (H 2 , He, NH 3 ), eliminating the need for interference correction equations.

Speciation Tools A wide variety of analytical and chemical techniques have be en used to characterize a metal’s speciation (as defined above) in various media (Hunt et al., 1992; Manceau et al., 1996, 2000a; Welter et al., 1999; Szulcz ewski et al., 1997; Isaure et . al., 2002; Lumsdon and Evans, 1995; Gupta and Chen, 1975; Ma and Uren, 1995; Ch arlatchka et al., 1997). Perhaps the most important factor that one must keep in mind in selecting a technique is that, when dealing with metal-contaminated media, one is most often look ing for the proverbial “needle in a haystack.” Therefore, the speciation technique must not onl y provide the information outlined above, but it

must also determine that information from a medi um that contains very little of the metal. As illustrated in Figure AX7-1.1.5, for a Pb-contam inated soil, less than 1% (modally) of a single species can be responsible for a bulk metal’s concentration above an action level. This factor is even more significant for other metals (i.e., As, Cd, or Hg) were action levels are often below 100 mg/kg. Of the techniques tested (physic ochemical, extractive, and theore tical), the tools that have been used most often to evaluate speciation for particle-bound metal include X-ray absorption spectroscopy (XAS), X-ray diffraction (XRD), particle induced X-ray emission (PIXE and μPIXE), electron probe microanalysis (EPMA) , secondary ion mass spectrometry (SIMS), X-ray photoelectron spectroscopy (XPS), sequential ex tractions, and single chemical extractions.

The tools that have been used most often to evaluate speciation for metal particles in solution include the following computer-based models : MINTEQL, REDEQL2, ECOSAT, MINTEQA2, HYDRAQL, PHREEQE, and WATEQ4F. Thes e tools are briefly described below.AX7-12 Figure AX7-1.1.5. Bulk lead versus single sp ecies modality. Shade d area represents typical soil-lead concentrations fo und at remediation sites in the western United States.

Particle-Bound Metal Direct Approaches Over the past decade, numerous advances in materials science have led to the development of a wide range in analytical tool s for the determination of metal concentration, bonding, and valance of individual particles on a scale that can be considered useful for the speciation of environmentally important material s (i.e., soils, wastes, sediments, and dust). This review will provide the reader with a brief description of these techniques, including their benefits, limitations (cost, avai lability, sample preparation, reso lution), and usability as well as references to current applications. Although mo st of these tools are scientifically sound and offer important information on the mechanistic u nderstanding of metal o ccurrence and behavior, only a few currently provide useful information on metal bioavailability at a “site” level. However, one may still find other techniques essential to a detailed characterization of a selectedAX7-13 material to describe the chemical or kinetic f actors controlling a metal’s release, transport, and/or exposure. X-Ray absorption Spectroscopy (XAS) . X-ray absorption spectroscopy (XAS) is a powerful technique using the tunable, monochr omatic (white light) X-rays produced by a synchrotron (2-4 GeV) to record oscillations in atomic absorption within a few 100 GeV of an element’s absorption edge. Spectra provide in formation on both chemical state and atomic structure. Measurements are theoretically availa ble for all elements and are not surface-sensitive nor sample-sensitive (i.e., gases, liquids, so lids, and amorphous mate rials are testable). High-energy spectra within 30 eV of the edge, termed XANES (X-ray absorption near edge structure spectroscopy (Fendorf et al., 1994; Maginn, 1998), are particularly suited for determination and quantification (10 to 100 ppm ) of metal in a part icular oxidation state (Szulczewski et al., 1997; Sha ffer et al., 2001; Dupont et al., 2002). The lower-energy spectra persist some 100 eV above the edge. These os cillations are termed EXAFS (extended X-ray absorption fine structure) and are more commonl y used for speciation analyses (Welter et al., 1999; Manceau et al., 1996, 2000a; Isaure et al., 2002). The main limitations to XAS techniques are (1 ) the lack of spatial resolution; (2) XAS techniques provide only a wei ghted average signal of struct ural configurations, providing information on the predominant form of the metal, while minor species, which may be more bioavailable, can be overlooked; (3) access to sync hrotrons is limited and the beam time required to conduct a site investigation would be prohibitive; (4) a la rge spectral library must be developed; (5) generally, poor fits to solution models are achieved when the compound list is large; and (6) high atomic number elements ha ve masking problems based on compound density. X-Ray Diffraction (XRD). In X-ray diffraction, a monochromatic Fe, Mo, Cr, Co, W, or Cu X-ray beam rotates about a finely powdered sample and is reflected off the interplanar spacings of all crystalline compounds in the sample, fulfilling Bragg’s law (n λ = 2dsin θ ). The identification of a species from this pattern is based upon the position of the lines (in terms of
θ or 2 θ ) and their intensities as recorded by an X-ray detector. Th e diffraction angle (2 θ ) is determined by the spacing between a particular set of atomic planes . Identification of the species is empirical, and current available da tabases contain more than 53,000 compounds. If a sample contains multiple compounds, interpretation becomes more difficult and computer-matching programs are essential. In some instances, by measuring the intensity of theAX7-14 diffraction lines and comparing them to standa rds, it is possible to quantitatively analyze crystalline mixtures; however, if the species is a hydrated form or has a preferred orientation, this method is only semiquantitative at best. Si nce this technique repr esents a bulk analysis, no particle size or lability informati on can be extracted from the patterns. Particle Induced X-Ray Emission (PIXE and μPIXE). Particle induced X-ray emission (PIXE) uses a beam, ~4 μm in diameter, of heav y charged particles (gener ally He) to irradiate the sample. The resulting characteristic X-rays are emitted and detected in a similar manner as XRF, using Si-Li detectors. Pa rticles generated from a small accel erator or cyclotron, with a potential of 2 to 4 MeV, are commonly used. Detection limits on the order of 1 mg/kg are achieved on thin-film samples. Disadvantages to its use for speciation include (1) only a small volume of material can be analyzed (1 to 2 mg/cm 2 ); (2) no particle size information is provided; (3) peak overlaps associated with Si-Li dete ctors limit identification of species; (4) limited availability; and (5) high cost. For a further review of PIXE analysis and applications, see Maenhaut (1987). Electron Probe Microanalysis (EPMA) . Electron probe microanalysis uses a finely focused (1 μm) electron beam (generated by an electron gun operating at a 2 to 30 kV accelerating voltage and pico/nanoamp currents) to produce a combination of characteristic X-rays for elemental quantifica tion along with secondary electrons and/or backscatter electrons for visual inspection of a sample. Elements fr om beryllium to uranium can be nondestructively analyzed at the 50-ppm level with limited samp le preparation. X-ray spectra can be rapidly acquired using either wavelengt h dispersive spectrometers (WDS) or energy dispersive spectrometers (EDS). With WDS, a set of diffracting crystals, of known d-spacing, revolve in tandem with a gas-filled proportional counter insi de the spectrometer housing so that Bragg’s law is satisfied and a particular wavelength can be focused. P hoton energy pulses reflecting off the crystal are

collected for an individual elemental line by the counter as a fi rst approximation to concentration. For quantitative analysis, these intensities ar e compared to those of known

standards and must be corrected for background, dead time, and elemental interactions (ZAF) (Goldstein et al., 1992). ZAF co rrection is in reference to th e three components of matrix effects: atomic number (Z), ab sorption (A), and fluorescence (F).AX7-15 With EDS, a single Si-Li crystal detector is used in conjunction with a multichannel analog-digital converter (ADC) to sort electrical pulses (with he ights approximately proportional to the quantum energy of the photon that gene rated them), producing a spectrum of energy (wavelength) versus counts. The net area under a particular peak (element al line) is proportional to its concentration in the sample. For quantitativ e analyses, corrections similar to WDS analysis must be performed. Although EDS detectors are more efficient than WDS, detection limits are significantly greater (~1000 ppm), because of elevated backgrounds and peak overlaps. For speciation analysis, the EDS system must NEVER be used as the primary detector, as numerous errors in species identification are often made. These are generally the result of higher-order X-ray line overlaps. This technique has been routinely used for site characterizations (Linton et al., 1980; Hunt et al., 1992; Camp, Dresser, and McKee (CDM), 1994; U.S. Environmental Protection Agency, 2002). Currently this technique offers the most complete data package on metal speciation than any of the other tools. The method is relative ly fast and inexpensive, available, and provides all of the required information for bi oavailability assessment s (i.e., particle size, species, lability, and sourcing). A number of limitations still need to be addressed including: (1) its inability to quickly isolate a statistically significant population of particles in soils with low bulk metal concentrations (<50 mg/kg), meaning that for some metals with low concentrations of concern (i.e., Cd, Mo, Sb, Se), this method may be less useful; (2) the more volatile metals (i.e., Hg, Tl) ar e often volatilized under the electron beam or lost during sample preparation. Secondary Ion Mass Spectrometry (SIMS). Secondary ion mass spectrometry (also known as ion microprobes or ion probes) is a well-known technique, primarily surface focused, that uses a 0.5 to 20 kV O, Ar, Ga, In, or Cs ion beam in bombarding (sputtering) the surface of a sample while emitting secondary ions that are detected by either quadrapole, tim e-of-flight (TOF), or magnetic sector mass spectrometers. Sensitivity is very high, in the ppb range for elements hydrogen to uranium, providing quantitative results on elemental or isotopic metals and organic compounds. With the advent of liquid metal (In and Ga) ion beam s, beam sizes of less than 1 μm are possible, although 20 μm is more commonly used.AX7-16 The major disadvantage of SIMS to species id entification is that each element or isotope must be tuned and analyzed sequentially. This makes the identification of a metal form highly time-consuming and, thus, the characterizati on of a multiphase medium impractical. X-Ray Photoelectron Spectroscopy (XPS). X-ray photoelectron spectroscopy or ESCA (electron spectroscopy for chemical analysis, as it was previously known) is a classical surface, 10 to 50 Å in depth, analytical technique for dete rmining qualitative elemental concentrations of elements greater than He in atomic number and provides limited structural and oxidation state information. In XPS, the high-energy (15 kV) electrons are typically produced from a dualanode (Al-Mg) X-ray tube. The ex citation or photoionization of atom s within the n ear surface of the specimen emit a spectrum of photoelectrons. The measured binding energy is characteristic

of the individual atom to which it was bound. Monochromatic sources are often employed to improve energy resolution, allowing one to infer oxidation states of elements or structure of compounds (organic and inorganic) by means of small chemical shifts in binding energies (Hercules, 1970). The major disadvantages of XPS for environmental speciation studies is its poor sensitivity, especially in complex matrices and its large, 100-200 μm , spatial resolution. The direct speciation techniques discusse d above are summarized in Table AX7-1.1.3.

Indirect Approaches A more indirect approach to speciation than the methods previously described include the functional or operational extracti on techniques that have been used extens ively over the years (Tessier et al., 1979; Tessier and Campbell, 198 8; Gupta and Chen, 1975). These methods use either a single or sequential ex traction procedure to release speci es associated with a particular metal within the media. Single chemical extr actions are generally used to determine the

bioavailable amount of metal in a functional class: water-sol uble, exchangeable, organically bound, Fe-Mn bound, or insoluble. In a similar approach, sequential extractions tr eat a sample with a succession of reagents intended to specifically dissolve different, less av ailable phases. Many of these techniques have been proposed, most of which are a variation on the classical method of Tessier et al. (1979), in which metal associated with exchang eable, carbonate-bound, Fe-Mn bound, organically bound, and residual species can be determined. Beckett (1989), Kheboian and Bauer (1987), and Foerstner (1987) provide excellent reviews on the use and abuse of extractions. TheseAX7-17Table AX7-1.1.3. Characteristicsfor Direct Speciation TechniquesTools Species Lability Species Particle Size Species Valance State Species Bonding Species Composition Species Abundance Element Specificity Isotopic Character Element Sensitivity Resolution Availability CostXRD No No No No No# No No No 3-4 vol% Bulk 1 $EMPA Yes Yes Yes+ No Yes Yes? B-U No*** 50 ppm 0.5-1 μm 2 $$SIMS No Yes No No Yes* Yes** Li-U Yes 1 ppb 10 μm 4 $$$XPS No No Yes Yes Yes* Yes** H-U No wt.% 100 μm 2 $$XAS No No Yes Yes Yes* Yes** He-U No ppb 2 μm 5 $$$$PIXIE No No No No Yes Yes** B-U No 10 ppm 4 μm 4 $$$$* Technique requires each element be tuned and standardized, requiring unreasonable time limits.** Techniques designed and tested only on simple systems. Multiple species require lengthy analytical times and data reduction.*** Limited when combined with ICP/MS/LA.# Identifies crystalline compounds andstoichiometric compositions only.? Technique has limitations based on particle counting statistics.+ Valance determined by charge balance of complete analyses.AX7-18 techniques can be useful in a study of metal uptake in plants, where transfer takes place predominately via a solution pha se. However, one must keep in mind that they are not “selective” in metal species, give no particle size information and, above all, these leachable

fractions have never been co rrelated to bioavailability. Solution Speciation Using Computer-Based Models. Computer-based models are either based upon equilibrium constant s or upon Gibb’s free energy valu es in determining metal speciation from solution chemistry conditions (c oncentration, pH, Eh, organic complexes, adsorption/desorption sites, and temperature). Bo th approaches are subject to mass balance and equilibrium conditions. These models have under gone a great deal of development in recent years, as reliable thermodynamic data has beco me available and can provide some predictive estimates of metal behavior. A good review of th ese models and their applications is provided by Lumsdon and Evans (1995). Speciation can be controlled by simple reacti ons; however, in many cases, particularly in contaminated media, their state of equilibrium and reversibility are unknown. In addition, these

models suffer from other limitations such as a lack of reliable thermodynamic data on relevant species, inadequacies in models to correct for high ionic streng th, reaction kinetics are poorly known, and complex reactions with co-precipitation/adsorption are not modeled. The first limitation is perhaps the most signif icant for contaminated media. For example, none of the models would predic t the common, anthropogenic, Pb phases, i.e., paint, solder, and slag.

AX7.1.1.4 Biotic Ligand Model The biotic ligand model (BLM) is an equilibr ium-based conceptual model that has been incorporated into regulatory agencies guidelines (i ncluding the EPA) to predict effects of metals primarily on aquatic biota and to aid in the unde rstanding of their intera ctions with biological surfaces (see Annex Section AX7.2.1.3).

Because of assumed similarities in mechanisms of toxicity between aquatic and terrestrial organisms, it is likely that the BLM approach as developed for the aquatic compartment may also be applicable to the terrestrial environment. Recent research has been directed toward extending the BLM to predict metal toxicity in soils (Steenbergen et al., 2005). Steenbergen et al. (2005) pointed out that, until recently, the BLM concept ha s not been applied to predict toxicity to soilAX7-19 organisms. The authors believe there may be two reasons for this. First, metal uptake routes in soils are generally more complex than those in water, because exposure via pore water and exposure via ingestion of soil particles may, in pr inciple, both be important. Second, it remains very difficult to univariately control the co mposition of the soil pore water and the metal concentrations in the pore water, due to re-e quilibration of the system following modification of any of the soil properties (inc luding addition of metal salts). Steenbergen et al. (2005) assessed acute copper toxi city to the earthworm Aporrectodea caliginosa using the BLM. To overcome the aforemen tioned problems inherent in soil toxicity tests they developed an artificia l flow-through exposure system consisting of an inert quartz sand matrix and a nutrient solution, of which the co mposition was univariately modified. Thus, the obstacles in employing the BLM to terrestrial ecosystems seem to be surmountable, and future research may provide useful information on Pb bioavailability and toxi city to terrestrial organisms.

AX7.1.1.5 Soil Amendments The removal of contaminated soil to mitigate exposure of terrestrial ecosystem components to Pb can often present both economic and logistic problems. Because of this,

recent studies have focused on in situ methodol ogies to lower soil-Pb RBA (Brown et al., 2003a,b). To date, the most common methods studi ed include the addition of soil amendments in an effort either to lower the solubility of the Pb form or to provide sorbtion sites for fixation of pore-water Pb. These amendments typically fa ll within the categories of phosphate, biosolid,

and Al/Fe/Mn-oxide amendments.

Phosphate Amendments Phosphate amendments have been studied exte nsively and, in some cases, offer the most promising results (Brown et al., 1999; Ryan et al., 2001; Cotter-Howells and Caporn, 1996; Hettiarachchi et al., 2001, 2003; Rabinowitz, 1993; Ya ng et al., 2001; Ma et al., 1995). Research in this area stems from early work by Nria gu (1973) and Cotter-Howe lls and Caporn (1996), who pointed out the very low solubili ties for many Pb-phosphates (Ksp ! 27 to ! 66), particularly chloropyromorphite [Pb 5 (PO 4 ) 3 Cl]. The quest to transform sol uble Pb mineralogical forms into chloropyromorthite continues to be the primary focus of most studies. Sources of phosphorousAX7-20 have included phosphoric acid (H 3 PO 4 ), triple-super phosphate (TSP) , phosphate rock, and/or hydroxyapatite (HA). Various studi es have combined one or more of these phosphorous sources with or without lime, iron, and/or manganese in an attempt to enhance amendment qualities.

Most amendments are formulated to cont ain between 0.5 and 1.0% phosphorous by weight. They are then either applied wet or dry and then mixed or left unmixed with the contaminated soil. Success of phosphate amendments has been variable, and the degree of success appears to depend on available phosphorous and the dissolu tion rate of the original Pb species. A number of potentially signi ficant problems associated with phosphate amendments have been recognized, including both phyto- and earthw orm toxicity (Ownby et al., 2005; Cao et al., 2002; and Rusek and Marshall, 2000). Both of these toxicities are primarily associated with very high applications of phosphorous and/or decreased soil pH. Indications of phytotoxicity are often balanced by studies such as Zhu et al. (200 4) that illustrate a 50 to 70% reduction in shootroot uptake of Pb in phosphate-amended soils. Additionally, the added phosphate poses the potential risk of eutrophication of nearby waterways from soil runoff. Finally, Pb-contaminated

soils from the extractive metals industry or agricultural sites often have elevated concentrations

of arsenic. It has been s hown (Impellitteri, 2005; Smith et al., 2002; Chaney and Ryan, 1994; and Ruby et al., 1994) that the addition of phosphate to such soils would enhance arsenic mobility (potentially moving arsenic down in to the groundwater) through competitive anion exchange. Some data (Lenoble et al., 2005) indi cate that if one coul d amend arsenic and Pb contaminated soils with iron(III) phosphate this problem can be mitigated, however the increased concentrations of both phosphate an d iron still exclude th e application when drinking water is an

issue.

Biosolid Amendments Historically, biosolids have been used in th e restoration of coal mines (Haering et al., 2000; Sopper, 1993). More recently, workers have demonstrated the feasib ility of their use in the restoration of mine tailings (Brown et al., 2003a), and ur ban soils (Brown et al., 2003b; Farfel et al., 2005). Mine tailings are inherently difficult to remedi ate in that they pose numerous obstacles to plant growth. They are most often (1) acidic; (2) high in metal content, thus prone to phytotoxicity; (3) very low in organic content; an d (4) deficient in macro- and micronutrients.AX7-21 Stabilization (i.e., the establishment of a vegetativ e cover) of these environments is essential to the control of metal exposure or migration from soil/dust and groundwater pathways. At Bunker Hill, ID, Brown et al. (2003b) de monstrated that a mixture of high nitrogen biosolids and wood pulp or ash, when surface applied at a rate of approximately 50 and 220 tons/ha, respectively, increased soil pH from 6.8 to approximately 8.0, increased plant biomass from 0.01 mg/ha to more than 3.4 tons/ha, and resulted in a heal thy plant cover within 2 years. Metal mobility was more difficult to eval uate. Plant concentrations of Zn and Cd were generally normal for the first 2 years of the study; however, Pb concentrations in vegetation dramatically increased two to three times in the first year. Additionally, macronutrients (Ca, K, and Mg) decreased in plant tissue. Urban soils, whether contaminated from smelting, paint, auto emissions, or industrial activity, are often contaminated with Pb (Agenc y for Toxic Substances and Disease Registry [ATSDR], 1988) and can be a sign ificant pathway to elevated child blood Pb levels (Angle et al., 1974). Typically, contaminated residential soils are replaced under Superfund rules. However, urban soils are less likely to be remediated un less a particular facility is identified as the contaminate source. Application of biosolids to such soils may be a cost-effective means for

individuals or communities to lower Pb RBAs. A field study by Farfel et al. (2005) using the commercia l biosolid ORGO found that, over a 1-year period, Pb in the dripline soils of one residence had reduced RBAs by ~60%. However, soils throughout the re mainder of the yard showed e ither no reduction in RBA or a slight increase. A more complex study was conducted by Brown et al. (2003a) on an urban dripline soil in the lab. The study used an asso rtment of locally derived biosolids (raw, ashed, high-Fe compost, and compost) with and without lime. All amendments were incubated with approximately 10% biosolids for a little more th an 30 days. In vitro and in vivo data both indicated a 3 to 54% reduction in Pb RBA, w ith the high-Fe compost providing the greatest

reduction. As with phosphate amendments, problems with biosolid application have also been

documented. Studies have shown that metal tran sport is significantly accelerated in soils amended with biosolids (Al-Wabel et al., 2002 ; McBride et al., 1997, 19 99; Lamy et al., 1993; Richards et al., 1998, 2000). Some of these stud ies indicate that metal concentrations in soil solutions up to 80 cm below the amended surface in creased by 3- to 20-fold in concentration upAX7-22 to 15 years after biosolid applic ation. The increase in metal tr ansport is likely the result of elevated dissolved organic carbon (DOC) in th e amended soil. Anodic stripping voltammetry has indicated that very low percentages (2 to 18 %) of the soluble metals are present as ionic or inorganic complexes (McBride, 1999; Al-Wabel et al., 2002).



### AX7.1.2 Distribution of Atmos pherically Delivered Lead in Terrestrial Ecosystems The 1986 Lead AQCD (U.S. Environmental Pr otection Agency, 1986a) contains only a

few minor sections that detail the speciation, distri bution, and behavior of atmospherically delivered Pb in terrestrial ecosystems. The docum ent concluded that the majority of Pb in the atmosphere at that time was from gasoline consum ption: of the 34,881 tons of Pb emitted to the atmosphere in 1984, 89% was from gasoline us e and minor amounts were from waste oil combustion, iron and steel production, and smelting. Lead in the atmosphere today, however, is not primarily from leaded-gasoline consump tion, but results largely from iron and steel foundaries, boilers and process he aters, the combustion of fossil fuels in automobiles, trucks, airplanes, and ships, and othe r industrial processes (Table 2-8, Polissar et al., 2001; Newhook et al., 2003). The emission source can determine the species of Pb that are delivered to terrestrial ecosystems. For example, Pb species emitted from automobile exhaust is dominated by particulate Pb halides and double salts w ith ammonium halides (e.g., PbBrCl, PbBrCl 2 NH 4 Cl),

while Pb emitted from smelters is dominated by Pb-sulfur species (Hab ibi, 1973). The halides from automobile exhaust break down rapidly in the atmosphere, possibly via reactions with atmospheric acids (Biggins and Harrison, 1979) . Lead phases in the atmosphere, and presumably the compounds delivered to the surf ace of the earth (i.e., to vegetation and soils), are suspected to be in the form of PbSO 4 , PbS, and PbO (Olson and Skogerboe, 1975; Clevenger et al., 1991; Utsunomiya et al., 2004). There are conflicting reports of how atmospherically derived Pb specifically behaves in surface soils. This disagreement may represent the natural variability of the biogeochemical

behavior of Pb in different terrest rial systems, the different Pb s ources, or it may be a function of the different analytical methods employed. The importance of humic and fulvic acids (Zimdahl and Skogerboe, 1977; Gamble et al., 1983) and h ydrous Mn- and Fe-oxides (Miller and McFee, 1983) for scavenging Pb in soils are discussed in some detail in the 1986 Lead AQCD. NriaguAX7-23 (1974) used thermodynamics to argue that Pb-orthophosphates (e.g., pyrom orphite) represented the most stable Pb phase in many soils and sedime nts. He further suggest ed that, because of the extremely low solubility of Pb-phosphate mine rals, Pb deposition could potentially reduce phosphorous availability. Olson and Skoger boe (1975) reported that solid-phase PbSO 4 dominated gasoline-derived Pb speciation in surface soils from Colorado, Missouri, and Chicago,

while Santillan-Medrano and Juri nak (1975) suggested that Pb(OH) 2 , Pb(PO 4 ) 2 , and PbCO 3 could regulate Pb speciation in soils. However, insoluble organic materi al can bind strongly to Pb and prevent many inorganic phases from ev er forming in soils (Zimdahl and Skogerboe, 1977). The vertical distribution and mobility of at mospheric Pb in soils was poorly documented prior to 1986. Chapter 6 of the 1986 AQCD cited a few references sugges ting that atmospheric Pb is retained in the upper 5 cm of soil (Reav es and Berrow, 1984). Te chniques using radiogenic Pb isotopes had been developed to discern betw een gasoline-derived Pb and natural, geogenic (native) Pb, but these techniques were mostly app lied to sediments (Shiraha ta et al., 1980) prior to the 1986 Lead AQCD. Without using these tec hniques, accurate determinations of the depthdistribution and potential migra tion velocities for atmospherically delivered Pb in soils were largely unavailable. Several technological advances , combined with the expans ion of existing technologies after 1986 resulted in the publica tion of a large body of literatu re detailing the speciation, distribution, and geochemical behavi or of gasoline-derived Pb in the terrestrial environment. Most notably, the development of selective chemical extraction (S CE) procedures as a rapid and inexpensive means for partitioning Pb into differe nt soil and sediment phases (e.g., Pb-oxides, Pb-humate, etc.) has been exploited by a numbe r of researchers (Tessi er et al., 1979; Johnson and Petras, 1998; Ho and Evans, 2000; Scheckel et al., 2003). Also, since 1986, several workers have exploited synchrotron-based XAS in order to probe the electron co ordination environment of Pb in soils, organic matter, organisms, a nd sediments (Manceau et al., 1996; Xia et al., 1997; Trivedi et al., 2003). X-ray absorp tion studies can be used for th e in-situ determ ination of the valence state of Pb and can be used to quantify Pb speciation in a variety of untreated samples. Biosensors, which are a relatively new technolog y coupling biological material, such as an enzyme, with a transducer, offer a new, simple , and inexpensive means for quantifying available Pb in ecosystems (Verma and Singh, 2005). Advan ces in voltammetric, diffusive gradients inAX7-24 thin films (DGT), and ICP techniques have also increased the abilities of researchers to quantify Pb phases in solutions (Berbel et al., 2001; Scally et al., 2003). In addition to the development of techniques for describing and quantifying Pb speci es in the soils and solutions, researchers have used radiogenic Pb isotopes ( 206 Pb, 207 Pb, 208 Pb) to quantify the dist ribution, speciation, and transport of anthropogenic Pb in soil profiles and in vegetation (Bindler et al., 1999; Erel et al., 2001; Kaste et al., 2003; Kl aminder et al., 2005). Over the past several decades, workers have also developed time-series data for Pb in precipitation, vegetation, organic horizons, mineral soils, and surface waters. Since atmospherically delivered Pb often comprises a si gnificant fraction of the “l abile” Pb (i.e., Pb not associated with primary minerals), these data have been useful for developing transport and

residence time models of Pb in different terrestrial re servoirs (Friedland et al., 1992; Miller and Friedland, 1994; Johnson et al., 1995a; Wang and Benoit, 1997). Overall, a significant amount of research has been publishe d on the distribution, sp eciation, and behavior of anthropogenic Pb in the terrestrial environment since 1986. However, certain specific details on the behavior of Pb in the terrestrial environment and its potential effects on soil microorganism s remain elusive.

AX7.1.2.1 Speciation of Atmospherically-del ivered Lead in Terrestrial Ecosystems

Lead in the Solid Phases Lead can enter terrestrial ecosystems through natural rock weatheri ng and by a variety of anthropogenic pathways. These different source terms control the species of Pb that is

introduced into the terrestrial environment. While Pb is highly concentrated (percent level) in certain hydrothermal sulfide deposits (e.g., PbS) that are disseminated throughout parts of the upper crust, these occurrences are relatively rare. Therefore, th e occurrence of Pb as a minor constituent of rocks (ppm level) , particularly granites, rhyolit es, and argillaceous sedimentary rocks is the more pertinent source term for the va st majority of terrestrial ecosystems. During the hydrolysis and oxidation of Pb-containing mine rals, divalent Pb is released to the soil solution where it is rapidly fixed by organic matter and se condary mineral phases (KabataPendias and Pendias, 1992). The geochemical form of natural Pb in terrestrial ecosystems will be strongly controlled by soil type (Emmanuel a nd Erel, 2002). In cont rast, anthropogenically introduced Pb has a variety of different geochemi cal forms, depending on the specific source. While Pb in soils from battery reclama tion areas can be in the form of PbSO 4 or PbSiO 3 , Pb inAX7-25 soils from shooting ranges and paint spills is commonly found as PbO and a variety of Pb carbonates (Vantelon et al., 2005; La perche et al., 1996; Manceau et al., 1996). Atmospherically delivered Pb resulting from fossil fuel combus tion is typically introduced into terrestrial ecosystems as Pb-sulfur compounds and Pb oxides (Olson and Skogerboe, 1975; Clevenger et al., 1991; Utsunomiya et al., 2004). After deposition, Pb species are likely transformed. Although the specific factors that control the speciation of anthr opogenic Pb speciation in soils are not well understood, there are many studies that have parti tioned Pb into its different geochemical phases. A thorough understanding of Pb speciation is critical in order to predict potential mobility and bioavail ability (see Section AX7.1.1).

Selective chemical extractions have been em ployed extensively over the past 20 years for quantifying amounts of a particular metal phase (e.g., PbS, Pb-humate, Pb-Fe/Mn-oxide) present

in soil or sediment rather than total metal concentration. Sometimes selective chemical extractions are applied sequentially to a particular sample. For example, the exchangeable metal

fraction is removed from the soil usin g a weak acid or salt solution (e.g., BaCl 2 ), followed immediately by an extraction ta rgeting organic matter (e.g., H 2 O 2 or NaOCl), further followed by an extraction targeting sec ondary iron oxides (e.g., NH 2 OH·HCl), and finally, a strong reagent cocktail (e.g., HNO 3 -HCl-HF) targets primary minerals. Te ssier et al. (1979) developed this technique. More recently, this technique has been modified and developed specifically for

different metals and different t ypes of materials (Keon et al., 2001). Alternatively, batch-style

selective chemical extractions have been used on soils and sediments to avoid the problems associated with nonselective reagents (Johnson and Petras, 1998). Selective extractions can be a relatively rapid, simple, and inexpensive means for determining metal phases in soils and sediments, and the generated data can be linked to potential mobility and bioavailability of the metal (Tessier and Campbell, 1987). However, some problems persist with the selective extraction technique. First, extrac tions are rarely specific to a single phase. For example, while

H 2 O 2 is often used to remove meta ls bound to organic matter in soils, others have demonstrated that this reagent destroys clay minerals and su lfides (Ryan et al., 2002). Peroxide solutions may also be inefficient in removing metals bound to humic acids, and in fact could potentially result in the precipitation of metal-humate substances. In addition to the nonselectivity of reagents, significant metal redistribution has been docum ented to occur during sequential chemical extractions (Ho and Evans, 2000; Sulkowski a nd Hirner, 2006), and many reagents may notAX7-26 completely extract targeted phases. While chemi cal extractions provide some useful information on metal phases in soil or sedime nt, the results should be treated as “operationally defined,” e.g.,

“H 2 O 2 -liberated Pb” rather than “organic Pb.” Lead forms strong coordination complexes wi th oxygen on mineral surfaces and organic matter functional groups (Abd-Elfattah and Wada , 1981), because of its high electronegativity and hydrolysis constant. Therefor e, Pb is generally not readily exchangeable, i.e., the amount of Pb removed from soils by dilute acid or salts is usually less that 10% (Karamanos et al., 1976; Sposito et al., 1982; Miller and McFee, 1983; Johnson and Petras, 1998; Bacon and Hewitt, 2005). Lead is typically adsorb ed to organic and inorganic so il particles str ongly via innersphere adsorption (Xia et al., 1997; Bargar et al., 1997a,b, 1998). Kaste et al. (2005) found that a single extract of 0.02 M HCl remove d 15% or less Pb in organic horizons from a montane forest in New Hampshire. The fact that relatively concentrated ac ids, reducing agents, oxidizing agents, or chelating agents are requ ired to liberate the majority of Pb from soils is used as one line of evidence that Pb migration and uptake by pl ants in soils is expected to be low. Lead that is “organically bound” in soils is typically quantified by extractions that dissolve/disperse or destroy organic matter. The former approach often employs an alkaline solution (NaOH), which deprotonates organic ma tter functional groups, or a phosphate solution,

which chelates structural cations. Extractions used to destroy organic matter often rely on H 2 O 2 or NaOCl. Both organic and mineral horizons typi cally have significant Pb in this soil phase. Miller and McFee (1983) used Na 4 P 2 O 7 to extract organically bound Pb from the upper 2.5 cm of soils sampled from northwestern Indiana. They found that organically bound Pb accounted for
between 25 and 50% of the total Pb present in the sampled topsoils . Jersak et al. (1997), Johnson and Petras (1998), and Kaste et al. (2005) selectively extracted Pb from spodosols from the

northeastern United States . Using acidified H 2 O 2 , Jersak et al. (1997) found that very little (<10 %) of the Pb in mineral soils (E, B, C) sampled from New York and Vermont was organic. Johnson and Petras (1998) used K 4 P 2 O 7 to quantify organically bound Pb in the Oa horizon and in mineral soils from the Hubbard Brook Experimental Forest in New Hampshire. They reported
that 60% of the total Pb in the Oa horizon was organic and that between 8 and 17% of the total Pb in the mineral soil was organic. However, in the E, Bh, and Bs1 horizons, organically bound Pb dominated the total “labile” ( non-mineral lattice) Pb. Kaste et al. (2005) used selective chemical extractions on organic horizons from m ontane forests in Vermont and New Hampshire.AX7-27 They found that repeated extractions with Na 4 P 2 O 7 removed between 60 and 100% of the Pb from their samples. Caution should be used when interpreting the results of pyrophosphate extractions. Although they are of ten used to quantify organically bound metals, this reagent can both disperse and dissolve Fe phases (Jeanroy and Guillet, 1981; Shuman, 1982). Acidified

H 2 O 2 has also been reported to destroy and rel ease elements associated with secondary soil minerals (Papp et al., 1991; Ryan et al., 2002). Aside from organic forms, Pb is often f ound to be associated with secondary oxide minerals in soils. Pb can be partitioned w ith secondary oxides by a variety of mechanisms, including (1) simple ion excha nge, (2) inner-sphere or outer -sphere adsorption, and (3) coprecipitation and/or occlusion (Bargar et al., 1997a,b, 1998, 1999). As discussed above, very

little Pb is removed from soil via dilute acid or salt solutions, so adsorption and co-precipitation are likely the dominant Pb interactions with secondary mineral phases. Reagents used to quantify this phase are often so lutions of EDTA, oxal ate, or hydroxylamine hydrochloride (HH). Miller and McFee (1983) used an EDTA solu tion followed by an HH solution to quantify Pb occluded by Fe and Mn minerals, respectively, in th eir surface-soil samples from Indiana. They reported that approximately 30% of the total soil Pb was occluded in Fe minerals, and 5 to 15% was occluded in Mn phases. In soils from the northeastern United States, Jersak et al. (1997) used various strengths of HH solutions and concl uded that negligible Pb was associated with
Mn-oxides and that 1 to 30% of th e Pb was associated with Fe phase s in the mineral soils in their study. Johnson and Petras (1998) reported that no Pb was remove d from the Oa horizon at the Hubbard Brook Experimental Forest (HBEF) by oxalate, but that 5 to 15% of the total Pb in mineral soils was removed by this extraction, presumably because it was bound to amorphous oxide minerals. Kaste et al. (2005), however, reported that HH removed 30 to 40% of the Pb from organic horizons in their study. They concluded that Fe phases were important in scavenging Pb, even in soil horizons dominated by organic matter. Synchrotron radiation (X-rays) allows researchers to probe the electron conf iguration of metals in untreated soil and sediment samples. Th is type of analysis has been extremely valuable for directly determining the coordi nation environment of Pb in a variety of soils and sediments.

Since different elements have di fferent electron binding energies (E b ), X-rays can be focused in an energy window specific to a metal of interest. In experiments involving XAS, X-ray energy is increased until a rapid increase in the amount of absorption occurs; this absorption edgeAX7-28 represents E b . The precise energy required to dislodge a core electron from a metal (i.e., E b ) will

be a function of the oxidation stat e and covalency of the metal. X-ray absorption studies that focus on the location of the absorption edge are referred to as XANES (X-ray absorption near edge structure). In the ener gy region immediately after the absorption edge, X-ray absorption increases and decreases with a periodicity that represents the wave f unctions of the ejected

electrons and the constructive and destructive inte rference with the wave functions of the nearby atoms. X-ray absorption studies used to inve stigate the periodicity of the absorption after E b are referred to as EXAFS (extended X-ray absorp tion fine structure). Since the electron configuration of a Pb atom will be directly governed by its speciation (e.g., Pb bound to organics, Pb adsorbed to oxide surfaces, PbS, etc.) X- ray absorption studies provide a powerful in-situ technique for determining specia tion without some of the proble ms associated with chemical extractions (Bargar et al., 1997a,b, 1998). Manceau et al. (1996) used EXAFS to study so il contaminated by gasoline-derived Pb in

France and found that the Pb was divalent a nd complexed to salicylate and catechol-type functional groups of humic substances. He conc luded that the alkyl-tet ravalent Pb compounds that were added to gasoline we re relatively unstable and will not dominate the speciation of Pb fallout from the combustion of leaded gasoline. The binding mechanism of Pb to organics is primarily inner-sphere adsorption (Xia et al., 1997). DeVolder et al. (2003) used EXAFS to demonstrate that Pb phases were shifting to th e relatively insoluble PbS when contaminated

wetland soils were treated with su lfate. Strawn and Sparks (2000) gave evidence that added Pb was adsorbed directly onto organic matter in an untreated silt loam, but in the absence of organic matter (treated with sodium hypochl orite) the added Pb appeared to adsorb instead to some form

of SiO 2 . More recent XAS studies have demonstrated the importance of biom ineralization of Pb in soils by bacteria and nematodes (Xia et al., 1997; Templeton et al., 2003a,b; Jackson et al., 2005). Templeton et al. (2003a,b) demonstrated that biogenic pr ecipitation of pyromorphite was the dominant source of Pb uptake by Burkholderia cepacia biofilms below pH 4.5. Above pH 4.5, adsorption complexes began to form in addition to Pb mineral precipitation.

In addition to XAS studies of Pb in environm ental samples, numerous experimental-based

XAS studies have documented in detail the coor dination environment of Pb adsorbed to Feoxides, Mn-oxides, Al-oxides, and clay mine rals (Manceau et al., 1996, 2000a,b, 2002; Bargar et al., 1997a,b, 1998, 1999; Strawn and Sparks, 1999; Trivedi et al., 2003; Chen et al., 2006).AX7-29 Bargar et al. (1997a) showed that Pb can adsorb to FeO 6 octahedra on three different types of sites: on corners, edges, or faces. Ostergre n et al. (2000a,b) showed that the presence of

dissolved carbonate and sulfate increased Pb adsorbtion on goethite. The relative fraction of corner-sharing complexes can be greatly increased by the presence of these ligands, as bridging complexes between the metal and the corners are formed (Ostergren et al., 2000a,b). Furthermore, Elzinga and Sparks showed th at the mechanism of Pb sorption to SiO 2 surfaces is pH-dependent. In acid environments (pH < 4) Pb adsorbs largely as an inner-sphere mononulear complex, but as pH increases, Pb sorption incr easingly occurs via the formation of surfaceattached covalent polynuclear Pb species. Recently, Jackson et al. (2005) used microfoc used synchrotron-based X-ray fluorescence ( Φ SXRF) to detail the distribution of Pb and Cu in the nematode Caenorhabditis elegans. They found that, while Cu was evenly distri buted throughout the bodies of exposed C. elegans, Pb was concentrated in the anterior pharynx region. Mi crofocused X-ray diffrac tion indicated that the highly concentrated Pb region in the pharynx was actually comp rised of the crystalline Pb mineral, pyromorphite. The authors concluded that C. elegans precipitated pyromorphite in the pharynx as a defense mechanism to prevent spre ading the toxic metal to the rest of the organism’s body. They further suggested that, becau se of the high turnover rate of nematodes, biomineralization could play an important role in the speciation of Pb in certain soils.

Lead Solid-solution Partitioning The concentration of Pb species dissolved in soil solution is probabl y controlled by some combination of (a) Pb-mineral solubility equili bria, (b) adsorption reactions of dissolved Pb phases on inorganic surfaces (e.g., crystalline or amorphous oxides of Al, Fe, Si, Mn, etc.; clay minerals), and (c) adsorption reactions of dissolv ed Pb phases on soil organic matter. Dissolved Pb phases in soil solution can be some combination of Pb 2+ and its hydrolysis species, Pb bound to dissolved organic matter, and Pb comp lexes with inorganic ligands such as Cl ! and SO 4 2 !

. Alkaline soils typically ha ve solutions supersaturat ed with respect to PbCO 3, Pb 3 (CO 3 ) 2 (OH) 2 , Pb(OH) 2 , Pb 3 (PO 4 ) 2 , Pb 5 (PO 4 ) 3 (OH), and Pb 4 O(PO 4 ) 2 (Badawy et al., 2002). Pb-phosphate minerals in particular are very insoluble, and calculations ba sed on thermodynamic data predict that these phases will control dissolved Pb in soil solution under a variety of conditions (Nriagu,AX7-30 1974; Ruby et al., 1994). However, certain chelat ing agents, such as dissolved organic matter, can prevent the precipitation of Pb minerals (Lang and Kaupenjohann, 2003). Using a combination of desorption experi ments and XAS, EXAFS, and XANES, Rouff et al. (2006) found that aging of Pb-calcite suspensions resulted in changes in the solid-phase distribution of Pb. Increased sorption time can reduce trace metal desorption by enhancing the stability of surface complexes (Rouff et al ., 2005) or by mechanisms involving microporous diffusion (Backes et al., 1995), re crystallization-induced incor poration into the solid phase (Ainsworth et al., 1994), and formation and stab ilization of surface precipitates (Ford et al., 1999). For adsorption of Pb to hydrous Fe oxides, goethite, and Pb-contaminated soils, aged samples show Pb to be reversibly bound, suggest ing Pb adsorption primarily to the substrates’ surfaces (Rouff et al., 2006). However, for Pb ad sorption to calcite aging played a significant role due to detection of multiple sorption mechan isms, even for short sorption times (Rouff et al., 2006). pH also played a role in Pb sorption. Over long sorption periods (60 to 270 days), slow continuous uptake of Pb occurred at pH 7.3 a nd 8.2. At pH 9.4, no further uptake occurred with aging and very little desorption occurred. These results show th e importance of contact time and pH on Pb solid-phase partitioning, particularly in geochemical systems in which calcite may be the predominant mineralogical constituent. Soil solution dissolved organic matter content and pH typically have very strong positive and negative correlations, respec tively, with the concentration of dissolved Pb species (Sauvé et al., 1998, 2000a, 2003; Weng et al., 2002; Badawy et al., 2002; Tipping et al., 2003). In the case of adsorption phenomena, the partitioning of Pb 2+ to the solid phase is also controlled by total metal loading, i.e., high Pb loadings will re sult in a lower fraction be ing partitioned to the solid phase. Sauvé et al. (1997, 1998) demonstr ated that only a fraction of the total Pb in solution was actually Pb 2+ in soils treated with leaf compost. The fraction of Pb 2+ to total dissolved Pb ranged from <1 to 60%, depending on pH and the availability of Pb-binding ligands. Nolan et al. (2003) used Donnan dialysis to show that 2.9 to 48.8% of the dissolved Pb

was Pb 2+ in pore waters of agricultural and contam inated soils from Australia and the United States. In acidic soils, Al sp ecies can compete for sites on natu ral organic matter and inhibit Pb binding to surfaces (Gustafsson et al., 2003). Differential pulse anodic stripp ing voltammetry (DPASV) is a t echnique that is useful for identifying relatively lo w concentrations of Pb 2+ and has found many applications in adsorptionAX7-31 and partitioning experiments. Th is technique has been particular ly useful for quantifying the K d , or partitioning ratio of Pb in the solid-to-liquid phase (K d = [total solid-phase metal in mg kg Γ 1 ] / [dissolved metal in mg L ! 1 ]). While the exact K d value is a function of pH, organic matter content, substrate type, total metal burden, and c oncentrations of competing ligands, such studies typically show that Pb has very strong solid-phase partitioning. Partitioning ratios determined by DPASV generally range from 10 3 to 10 6 in soils in the typical pH range (Sauvé et al., 2000b). Aualiitia and Pickering (1987) used thin film ASV to compare th e relative affinity of Pb for different inorganic particulates. They reported that Mn(IV) oxides completely adsorbed the Pb, regardless of pH in the range of 3 to 9, and had the highest affinity for Pb in their study. The adsorption of Pb to pedogenic Fe-oxides, Al-hydroxi des, clay minerals, and Fe ores was reported to be pH-dependent. Sauvé et al. (1998) used DPASV to study the effects of organic matter and pH on Pb adsorption in an orchard soil. They demonstrated that Pb complexation to dissolved organic matter (DOM) increased Pb solubility, and that 30 to 50% of the dissolved Pb was bound to DOM at pH 3 to 4, while >80% of the dissolved Pb was bound to DOM at neutral pH. They concluded that in most soils, Pb in solution would not be found as Pb 2+ but as bound to DOM. Sauvé et al. (2000a) compared the relative affinity of Pb 2+ for synthetic ferrihydrite, leaf compost, and secondary oxide minerals collected from soils. They reported that the inorganic mineral phases were more efficient at lowering the amount of Pb 2+ that was available in solution than the leaf compost. Glover et al. (2002) used DPASV in studying th e effects of time and organic acids on Pb adsorption to goethite. They found that Pb adsorption to geothite was very rapid, and remained unchanged after a period of about 4 h. Lead desorption was found to be much slower. The presence of salicylate appeared to increase the amount of Pb that desorbed from goethite more so than oxalate.

AX7.1.2.2 Tracing the Fate of Atmospherically-d elivered Lead in Terrestrial Ecosystems Radiogenic Pb isotopes offer a powerful tool for separating anthropogenic Pb from natural Pb derived from mineral weathering (Erel and Pa tterson, 1994; Erel et al ., 1997; Semlali et al., 2001). This has been particularly useful for studying Pb in mineral soil, where geogenic Pb

often dominates. The three ra diogenic stable Pb isotopes ( 206 Pb, 207 Pb, and 208 Pb) have a heterogeneous distribution in the earth’s crust pr imarily because of the differences in the halflives of their respective parents ( 238 U, T 1/2 = 4.7 × 10 9 year; 235 U, T 1/2 = 0.7 × 10 9 year; 232 Th,AX7-32 T 1/2 = 14 × 10 9 year). The result is that the ore bodies from which anthropogenic Pb are typically derived are usually enriched in 207 Pb relative to 206 Pb and 208 Pb when compared with Pb found in granitic rocks. Graney et al. (1995) analyzed a dated core from Lake Erie, and found that the

206 Pb/ 207 Pb value in sediment deposited in th e late 1700s was 1.224, but in 20th-century sediment, the ratio ranged from 1.223 to 1.197. This shift in the Pb isotopic composition represents the introduction of a significant amount of anthropogenic Pb into the environment. Bindler et al. (1999) and Emmanue l and Erel (2002) analyzed the isotopic composition of Pb in soil profiles in Sweden and the Czech Republic, respectively, and determined that mineral soils immediately below the organic horizon had a mixt ure of both anthropogenic and geogenic Pb. Anthropogenic Pb has been detect ed relatively deep in to the soil profile (>25 cm) in Europe, presumably this represents very “old”( i.e., pre-1900) Pb (Steinnes and Friedland, 2005). Erel and Patterson (1994) used radiogenic Pb isotopes to trace the movement of industrial Pb from topsoils to groundwaters to streams in a remote mountainous region of Yosemite National Park in California. They calculated th at total 20th-century indus trial Pb input to their study site was approximately 0.4 g Pb m Γ 2 . Lead concentrations in or ganic material were highest in the upper soil horizons, and d ecreased with depth. During snowmelt, Pb in the snowpack was mixed with the anthropogenic and geogenic Pb alr eady in the topsoil, a nd spring melts contained a mixture of anthropogenic and geogenic particul ate Pb. During base flows, however, 80% of the Pb export from groundwater and streams was from natural granite weathering (Erel and Patterson, 1994).

Uranium-238 series 210 Pb also provides a tool for traci ng atmospherically delivered Pb in

soils. After 222 Rn (T 1/2 = 3.8 days) is produced from the decay of 226 Ra (T 1/2 = 1600 years), some

fraction of the 222 Rn escapes from rocks and soils to the atmosphere. It then decays relatively rapidly to 210 Pb (T 1/2 = 22.3 years), which has a troposphe ric residence time of a few weeks (Koch et al., 1996). Fallout 210 Pb is deposited onto forests via wet and dry deposition, similar to anthropogenic Pb deposition in forests, and is thus useful as a tracer for non-native Pb in soils. Lead-210 is convenient to use for calculating the re sidence time of Pb in soil layers, because its atmospheric and soil fluxes can be assumed to be in steady state at undisturbed sites (Dörr and Münnich, 1989; Dörr, 1995; Kast e et al., 2003). Atmospheric 210 Pb ( 210 Pb ex hereafter, 210 Pb in “excess” of that supported by 222 Rn in the soil) must be calcul ated by subtracting the amount ofAX7-33 210 Pb formed in soils by the in-situ decay of 222 Rn from the total 210 Pb (Moore and Poet, 1976; Nozaki et al., 1978). Benninger et al. (1975) measured fallout 210 Pb in soils and stream water at Hubbard Brook

and at an undisturbed forest in Penns ylvania. They estimated atmospheric 210 Pb export in streamwaters to be <0.02% of the standing 210 Pb crop in the organic horizons. They used a simple steady-state model to calculate the reside nce time of Pb in the organic horizons to be 5,000 years. This overestimate of the Pb reside nce time in the organic horizons was likely a result of the low resolution of their sampling. Since they only sampled the upper 6 cm of soil and the drainage waters, they did not accurately evaluate the distribution of 210 Pb in the soil column in between. Dörr and Münnich (1989, 1991) used 210 Pb profiles in soils of southern Germany to evaluate the behavior of atmospherically delivered P b. They calculated the vertical velocity of Pb by dividing the rela xation depth (i.e., th e depth at which 210 Pb activity decreases to

1/ e , or approximately 37% of its surface value) by the 210 Pb mean life of 32 years. They reported downward transit velocities of atmospheri cally deposited Pb at 0.89 ± 0.33 mm year Γ 1 . The downward transport of atmospheric Pb was not aff ected by pH or soil type. However, since Pb velocities in the soil profile where identic al to carbon velociti es calculated with 14 C, they

concluded that Pb movement in forest soils is probably controlled by carbon transport. Kaste et al. (2003) used 210 Pb to model the response time of atmo spherically delive red Pb in the O horizon at Camel’s Hump Mountain in Vermont. They concluded th at the forest floor response time was between 60 and 150 years, depending on vegetation zone and elevation. Using

206 Pb: 207 Pb, they also demonstrated that some gaso line-derived Pb migrated out of the O horizon

and into the mineral soil in th e deciduous vegetation zone on the mountain, while all of the atmospheric Pb was retained in the upper 20 cm of the soil profile. Klaminder et al. (2006) used

210 Pb distributions to calculate the residence time of Pb in soils of Northern Sweeden. They concluded that atmospheric Pb ha d a residence time of ~200 year s in the mor layer of mature forests. They also provided evidence suggestin g that forests in an ea rlier stage of ecological

succession have a much shorter Pb reside nce time in the mor horizon (<50 years). Researchers assessing the fate of atmospheric Pb in soils have also relied on repeated sampling of soils and vegetation for total Pb (e.g., Zhang, 2003). This technique works best when anthropogenic Pb accounts for the vast majority of total Pb in a particular reservoir. Johnson et al. (1995a), Yanai et al. (2004), and Friedland et al. (1992) used O horizon (forestAX7-34 floor) time-series data to evaluate the movement of gasoline-derived Pb in the soil profile. These

studies have concluded that the distribution of Pb in the upper soil horizons has changed over the past few decades. Yanai et al . (2004) documented a decline in Pb from the Oie horizon between the late 1970s to the early 1990s in remote forest soils in New Hampshire. Johnson et al. (1995a) and Friedland et al. (1992) dem onstrated that some fraction of Pb had moved from the O horizon to the mineral soil during the 1980s at Hubbard Brook and at selected remote sites in the northeastern United States, respectively. Evans et al. (2005) demonstrated th at Pb concentrations in the litter layer (fresh litte r + Oi horizon) sampled in a transect from Vermont to Quebec decreased significantly between 1979 and 1996, reflec ting a decrease in Pb deposition to forests and upper soil horizons during that time period. Miller et al. (1993) and Wang and Benoit (1997) used forest floor time-series data to model the response time ( e folding time, the time it

takes a reservoir to decrease to the 1/ e , (ca. 37%) of its original amount ) of Pb in the forest floor. Miller et al. (1993) calculated O horizon response times of 17 years for the northern hardwood forest and 77 years in the spruce-fir zone on Camel’s Hump Mountain in Vermont. Wang and Benoit (1997) determined that the O horizon wo uld reach steady state with respect to Pb (1.3 μg g Γ 1 Pb) by 2100. Both suggested that the vert ical movement of organic particles dominated Pb transport in the soil profile.

AX7.1.2.3 Inputs/Outputs of Atmospherically-d elivered Lead in Terrestrial Ecosystems The concentration of Pb in contemporary rain fall in the mid-Atlantic and northeastern United States is on the order of 0.5 to 1 μg/L (Wang et al., 1995; Kim et al., 2000; Scudlark et al., 2005). Long-term trends in nine elementa l concentrations (includi ng Pb) in wet deposition

were measured on the Delmarva Peninsula near Lewes, Delaware, over the period from 1982 to 1989 (Figure AX7-1.2.1) (Scudlark et al., 1994). Of the nine elements measured, only Pb (and possibly Al and Cd), indicated a decreasing trend over time (3 μg/L in 1982 to <1.0 μg/L in 1989). The authors attributed the dramatic decr ease in Pb to the phasing out of Pb-alkyl additives from gasoline as mandated in the 1972 Clean Air Act. For comparison, rainfall measured in Los Angeles (CA) during 2003-2004 av eraged 0.15 μg/L, but showed nearly an order-of-magnitude variation, presumably because of the arid environmental (Sabin et al., 2005).

See also Table 2-22 for additional rainfall information.AX7-35 Figure AX7-1.2.1. Long-term (1982-1989) annual trends in lead concentrations (μg/L) in Lewes, Delaware precipitation.

Source: Scudlark et al. (1994)

The role of dry deposition in the total deposition of Pb to terrestrial ecosystems is not

constrained well. Researchers have estimat ed that dry deposition accounts for anywhere
between 10 to >90% of total Pb deposition (Galloway et al., 1982; Wu et al., 1994; Sabin et al., 2005). Migon et al. (1997) found total, we t, and dry Pb deposition to be 8.6, 1.6 and 7.0 μg/m2/d, respectively, over the Ligurian Sea, France. Dry deposition velocities were estimated at approximately 0.2 cm/sec, similar to anthropogenic element deposition velocities measured at remote lakes around Lake Michigan (Y i et al., 2001). Yi et al. (2001) found that dry deposition velocities for elements were correlated with particle concentrations in the following order: coarse > total particle > fine particle. Refer to Table 2-21 for additional information on dry, wet, and total Pb deposition velocities and fluxes. Arid environments appear to have a much higher fraction of dr y deposition: total deposition (Sabin et al., 2005). Furthermore, it is possible that Clean Air Act Legislation enacted in the late 1970s preferentially reduced Pb associated with fine particles, so the relativeAX7-36 contributions of dry deposition may have changed in the last few decades. If the major source of Pb to a terrestrial ecosystem is resuspended part iculates from transportation corridors, then the particle size fraction that domi nates deposition may be relatively coarse (>50 μm) relative to other atmospheric sources (Pirrone et al., 1995; Sansal one et al., 2003). Total contemporary loadings to terrestrial ecosystems are approximately 1 to 2 mg m Γ 2

year Γ 1 (Wu et al., 1994; Wang et al., 1995; Simonetti et al., 2000; Sabi n et al., 2005). This is a relatively small annual flux of Pb if compared to the reservoir of approximately 0.5 to 4 g m Γ 2 of gasoline-derived Pb that is already in surface soils over much of the United States (Friedland et al., 1992; Miller and Friedland, 1994; Erel and Patterson, 1994; Marsh and Siccama, 1997; Yanai et al., 2004; Johnson et al., 2004; Evans et al., 2005). While vegetation can play an important role in sequestering Pb from rain and dry deposition (Russell et al., 1981), direct uptake of Pb from soils by plants appears to be low (Klaminder et al ., 2005). High elevation areas, particularly those near th e base level of clouds often have higher burdens of atmospheric contaminants (Siccama, 1974). A Pb deposition m odel by Miller and Friedland (1994) predicted 2.2 and 3.5 g Pb m Γ 2 deposition for the 20th century in the deciduous zone (600 m) and the coniferous zone (1000 m), respec tively. More recently, Kaste etal. (2003) used radiogenic isotope measurements on the same mountain to c onfirm higher loadings at higher elevation. They measured 1.3 and 3.4 g gasoline-derived Pb m Γ 2 in the deciduous zone and coniferous zones, respectively. Higher atmospheric Pb load ings to higher elevatio ns are attributed to (1) the higher leaf area of coniferous species, which are generally more prevalent at high elevation; (2) higher rain fall at higher elevation; and (3) in creased cloudwater impaction at high elevation (Miller et al., 1993). Although inputs of Pb to ecosystems are curren tly low, Pb export from watersheds via groundwater and streams is substantially lower than inputs. Therefore, even at current input levels, watersheds are accumulating industrial Pb . However, burial/movement of lead over time down into lower soil/sediment layers also tends to sequester it away fr om more biologically active parts of the watershed (unless later di sturbed or redistribute d, e.g., by flooding, dredging, etc.). Seeps and streams at the HBEF have Pb concentrations on the order of 10 to 30 pg Pb g Γ 1 (Wang et al., 1995). At a remote valley in the Si erra Nevada, Pb concentrations in streamwaters were on the order of 15 pg Pb g Γ 1 (Erel and Patterson, 1994). Losses of Pb from soil horizons are assumed to be via particulates (Dörr and Münnich, 1989; Wang and Benoit, 1996, 1997). TylerAX7-37 (1981) noted that Pb losses from a spruce fore st A-horizon soil in Sweden were influenced by season; with highest Pb fluxes being observed during warm, wet months. He suggested that DOC production and Pb moveme nt were tightly linked.

Surface soils across the United States are enriched in Pb relative to levels expected from solely natural geogenic inputs (Friedland et al., 1984; Herrick and Friedland, 1990; Francek, 1992; Erel and Patterson, 1994; Ma rsh and Siccama, 1997; Yanai et al., 2004; Murray et al., 2004). While some of this contaminant Pb is attr ibuted to paint, salvage yards, shooting ranges,

and the use of Pb arsenate as a pesticide in localized areas (Fra ncek, 1997), Pb contamination of surface soils is essentially ubiquitous because of atmospheric pollution associated with metal smelting and production, the combustion of fossil fuels, and waste incine ration (Newhook et al., 2003; Polissar et al., 2001). Surface soils in Mich igan, for example, typically range from 8 to several hundred ppm Pb (Francek, 1992; Murray et al ., 2004). Soils collected and analyzed beneath 50 cm in Michigan, however, range only from 4 to 60 ppm Pb (Murray et al., 2004).

In remote surface soils from the Sierra Nevada Mountains, litter and uppe r soil horizons are 20 to 40 ppm Pb, and approximately 75% of this Pb ha s been attributed to atmospheric deposition during the 20th century (Erel and Patterson, 1994 ). Repeated sampling of the forest floor (O horizon) in the northeastern United States demonstrates that the organic layer has retained much of the Pb load deposited on ecosytems dur ing the 20th century. Total Pb deposition during the 20th century has been estimated at 1 to 3 g Pb m Γ 2 , depending on elevation and proximity to urban areas (Miller and Friedland, 1994; Johnson et al., 1995a). Forest floors sampled during the 1980s and 1990s, and in early 2000 had between 0.7 and 2 g Pb m Γ 2 (Friedland et al., 1992; Miller and Friedland, 1994; Johns on et al., 1995a; Kaste et al., 2003; Yanai et al., 2004; Evans et al., 2005). The pool of Pb in above- and belo w-ground biomass at the HBEF is on the order of 0.13 g Pb m Γ 2 (Johnson et al., 1995a).

The amount of Pb that has leached into mine ral soil appears to be on the order of 20 to 50% of the total anthropogenic Pb deposition. Kaste et al. (2003) and Miller and Friedland (1994) demonstrated that Pb loss from the forest floor at Camel’s Hump Mountain in Vermont depended on elevation. While the mineral soil in the deciduous forest had between 0.4 and 0.5 g

Pb m Γ 2 (out of 1 to 2 g Pb m Γ 2 in the total soil profile), at high er elevations the thicker coniferous

forest floor retained more than 90% of the tota l Pb deposition (Kaste et al., 2003). Johnson et al. (1995a) determined that the forest floor at HBEF in the mid-1980s had about 0.75 g Pb m Γ 2

.AX7-38 Compared to their estimated 20th-century atmospheric Pb deposition of 0.9 g Pb m Γ 2 , the forest floor has retained 83% of the atmospheric Pb loadings (Johnson et al., 1995a). Johnson et al. (2004) noted that gasoline-derived Pb was a signif icant component of the labile Pb at the HBEF. They calculated that Pb fluxes to the HBEF by at mospheric pollution were essentially equivalent to the Pb released by mineral weathering over the past 12,000 years. Marsh and Siccama (1997) used the relatively homogenous mineral soils und erneath formerly plowed land in Rhode Island, New Hampshire and Connecticut to assess the de pth-distribution of at mospheric Pb. They reported that 65% of the atmospheric Pb deposite d during the 20th century is in the mineral soil
and 35% is in the forest floor. At their remote study site in the Sierra Nevada Mountains, Erel and Patterson (1994) reported that most of the anthropogenic Pb was associated with the humus fraction of the litter layer and soils sampled in the upper few cm. Atmospherically delivered Pb is probably pres ent in ecosystems in a variety of different biogeochemical phases. A combination of Pb ad sorbtion processes and th e precipitation of Pb minerals will typically keep dissolved Pb sp ecies low in soil solution, surface waters, and streams (Sauvé et al., 2000a; Jackson et al., 2005) . While experimental and theoretical evidence suggest that the precipitation of inorganic Pb phases and the adsorption of Pb on inorganic phases can control the biogeochemistry of c ontaminant Pb (Nriagu, 1974; Ruby et al., 1994; Jackson et al., 2005), the influence of organic matt er on the biogeochemistry of Pb in terrestrial ecosystems cannot be ignored in many systems. Organic matter can bind to Pb, preventing Pb migration and the precipitation of inorganic phas es (Manceau et al., 1996 ; Xia et al., 1997; Lang and Kaupenjohann, 2003). As the abundance of organic matter declines in so il, Pb adsorption to inorganic soil minerals and the direct pr ecipitation of Pb phase s may dominate the biogeochemistry of Pb in terrestrial ecosystems (Ostergren et al., 2000a,b ; Sauvé et al., 2000a).

Conclusions Advances in technology since the 1986 Lead AQCD have allowed for a quantitative determination of the mobility, di stribution, uptake, and fluxes of at mospherically delivered Pb in ecosystems. Among other things, these studies have shown that industrial Pb represents a significant fraction of total labile Pb in wate rsheds. Selective chemical extractions and synchrotron-based X-ray studies have shown that industrial Pb can be strongly sequestered by organic matter and by secondary minerals such as clays and oxides of Al, Fe, and Mn. Some ofAX7-39 these studies have provided compelling evidence th at the biomineralizati on of Pb phosphates by soil organisms can play an important role in th e biogeochemistry of Pb. Surface soils sampled relatively recently demonstrate that the upper soil horizons (O + A horizons) are retaining most of the industrial Pb burden introduced to the sy stems during the 20th century. The migration and biological uptake of Pb in ecosystems is relativel y low. The different biogeochemical behaviors of Pb reported by various studies may be a resu lt of the many different analytical techniques employed, or they may be a result of natural vari ability in the behavior of Pb in different systems.

Lead Uptake into Plants Plants take up Pb via their foliage and th rough their root systems (U.S. Environmental Protection Agency, 1986a; Påhlsson, 1989). Surface deposition of Pb onto plants may represent a significant contribution to the total Pb in and on the plant, as has been observed for plants near smelters and along roadsides (U.S. Environmen tal Protection Agency, 1986a). The importance of atmospheric deposition on above-ground plant Pb uptake is well-documented (Dalenberg and Van Driel, 1990; Jones and Johnston, 1991; Ange lova et al., 2004). Data examined from experimental grassland plots in southeast England demonstrated th at atmospheric Pb is a greater contributor than soil-derived Pb in crop plan ts and grasses (Jones and Johnston, 1991). A study by Dalenberg and Van Driel (1990) showed that 75 to 95% of the Pb found in field-grown test plants (i.e., the leafy material of grass, spinach, and carrot; wheat grain; and straw) was from atmospheric deposition. Angelova et al. (2004) found that tobacco grown in an industrial area accumulated significant amounts of Pb from the at mosphere, although uptake from soil was also

observed. The concentration of Pb in tobacco se eds was linearly related to the concentration of Pb in the exchangeable and carbonate-bound fracti ons of soil, as measured using sequential extraction (Angelova et al ., 2004). Lead in soil is more signi ficant when considering uptake into root vegetables (e.g., carrot, potato), since, as was noted in th e 1986 Lead AQCD (U.S. Environmental Protection Agency, 1986a), most Pb remains in the roots of plants. There are two possible mechanisms (symplastic or apoplastic) by which Pb may enter the root of a plant. The symplastic route is thr ough the cell membranes of ro ot hairs; this is the mechanism of uptake for water and nutrients. Th e apoplastic route is an extracellular route between epidermal cells into the intercellular sp aces of the root cortex. Previously, Pb wasAX7-40 thought to enter the plant via the symplastic route, probably by tr ansport mechanisms similar to those involved in the uptake of calcium or other divalent cations (i.e., transpirational mass flow, diffusion, or active transport). However, it also had been speculated that Pb may enter the plant via the apoplastic route (U.S . Environmental Protection Agency, 1986a). Sieghardt (1990) determined that the mechanism of Pb uptake wa s via the symplastic route only and that the apoplastic pathway of transport was stopped in the primary roots by the endodermis. He studied the uptake of Pb into two plants, Minuartia verna (moss sandwort) and Silene vulgaris (bladder campion) that colonize metal-contaminated sites. In the roots of both plants, Pb was found mainly in the root cortex. Active ion uptake was required to transport the Pb into the stele and

then into the shoots of th e plant (Sieghardt, 1990). Although some plants translocate more Pb to the shoots than others, most Pb remains in the roots of plants. Two mechanisms have been proposed to account for this relative lack of translocation to the shoots: (1) Pb may be deposited within root cell wall material, or (2) Pb may be sequestered within root cell organelles (U.S. Environmental Protection Agency, 1986a). Påhlsson (1989) noted that plants can accumulate large quantities of Pb from the soil but that translocation to shoots and leaves is limited by th e binding of Pb ions at root surfaces and cell walls. In a study by Wierzbicka (1999), 21 di fferent plant species were exposed to Pb 2+ in the form of Pb-chloride. The pl ant species included cucumber ( Cucumis sativus ), soy bean ( Soja hispida ), bean ( Phaseolus vulgaris ), rapeseed ( Brassica napus ), rye ( Secale cereale ), barley ( Hordeum vulgare ), wheat ( Triticum vulgare ), radish ( Raphanus sativus ), pea ( Pisum sativum ), maize ( Zea mays ), onion ( Allium cepa ), lupine ( Lupinus luteus ), bladder campion ( Silene vulgaris ), Buckler mustard ( Biscutella laevigata ), and rough hawkbit ( Leontodon hispidus ). Although, the amount of Pb taken up by the plant varied with species , over 90% of absorbed Pb

was retained in the root s. Only a small amount of Pb was tr anslocated (~2 to 4%) to the shoots of the plants. Lead in roots was present in the deeper layers of r oot tissues (in particular, the root cortex) and not only on the root surface. Th ere was no correlation between Pb tolerance (measured as root mass increase expressed as a per centage of controls) and either root or shoot tissue concentrations (Wierzbicka, 1999). The study by Wierzbicka (1999) was the first to report that plants developing from bulbs, in this case th e onion, were more tolerant to Pb than plants developing from seeds. This tolerance was assume d to be related to the large amounts of Pb that were transported from the roots and stored in the bulb of the plant (Wierzbicka, 1999).AX7-41 (AX7-1) Uptake of Pb from soil into plants was m odeled as part of Eco-SSL development (U.S. Environmental Protection Agency, 2005a). The re lationship derived between Pb in the soil and Pb in a plant was taken from Bechtel Jac obs Company (BJC) (1998) and is as follows:

Ln(Cp) = 0.561 * Ln(Csoil)

− 1.328

where Cp is the concentration of Pb in the plant (dry weight) and Csoil is the concentration of Pb in the soil. This equation r ecognizes that the rati o of Pb concentration in plant to Pb

concentration in soil is not constant.

Invertebrates There was no clear evidence suggesting a differe ntial uptake of Pb into different species of earthworm ( Lumbricus terrestris , Aporrectodea rosea , and A. caliginosa ) collected around a smelter site near Avonmouth, Englan d (Spurgeon and Hopkin, 1996a). Th is is in contrast to Pižl and Josens (1995) and Terhivuo et al. (1994) who found Aporrectodea spp. accumulated more

Pb than Lumbricus . The authors suggested that these differences c ould be due to different feeding behaviors, as Lumbricus feeds on organic material and Apporectodea species are geophagus, ingesting large amounts of soil during f eeding. The differences between species also may be related to differing efficiencies in ex cretory mechanisms (Pižl and Josens, 1995). However, the interpretation of species differe nce is complicated by a number of potentially confounding variables, such as soil characteri stics (e.g., calcium or other nutrient levels) (Pižl and Josens, 1995).

The bioaccumulation of Pb from contaminat ed soil was tested using the earthworm Eisenia fetida , and the amount of Pb accumulated did not change significantly until the concentration within soil reached 5000 mg/kg (D avies et al., 2003). This coincided with the

lowest soil concentrations at which earthw orm mortality was observed. The ratio of the

concentration of Pb in worms to the concentr ation in soil decreased from 0.03 at 100 mg/kg to 0.001 at 3000 mg/kg, but then increased quickly to 0.02 at 5000 mg/kg. The authors concluded that earthworms exhibit regulated uptake of Pb at levels of low cont amination (<3000 mg/kg) until a critical concentration is reached, at whic h point this mechanism breaks down, resulting in unregulated accumulation and mortality. This study was conducted using test methods whereAX7-42 (AX7-2) soil was not allowed to equilibrate following the a ddition of Pb and prior to the addition of the test organisms. This may have resulted in an increased bioavailability and overestimated Pb toxicity relative to actual environmental condit ions (Davies et al., 2003). See the discussion in Section AX7.1.2 on the effects of aging on Pb sorption processes. Lock and Janssen (2002) and Bongers et al. ( 2004) found that Pb-nit rate was more toxic

than Pb-chloride to survival and reproduction of the springtail Folsomia candida . However, percolation (removal of the chlori de or nitrate counterion) causeda significant decrease in Pbnitrate toxicity such that there was no differe nce in toxicity once th e counterion was removed (Bongers et al., 2004). No change in toxicity was observed for Pb-chloride once the chloride was removed from the soil. Bongers et al. (2004) suggested that the nitr ate ion was more toxic than the chloride ion to springtails. Uptake of Pb from soil into earthworms was also modeled as part of Eco-SSL development (U.S. Environmental Protection Agen cy, 2005a). The relationship derived between

Pb in the soil and Pb in an earthworm was take n from Sample et al. ( 1999) and is as follows:

Ln(Cworm) = 0.807 * Ln(Csoil) – 0.218

where Cworm is the concentration of Pb in the earthworm (dry wei ght) and Csoil is the

concentration of Pb in the soil. This equation recognizes that th e ratio of Pb concentration in

worm to Pb concentration in soil is not constant.

Wildlife

Research has been conducted to determine wh at Pb concentrations in various organs

would be indicative of various le vels of effects. For example, Franson (1996) compiled data to

determine what residue levels were consistent with three levels of effects in Falconiformes (e.g., falcons, hawks, eagles, kestrels, ospreys), Co lumbiformes (e.g., doves, pigeons), and Galliformes (e.g., turkey, pheasant, partridge, quail, chickens). The three levels of eff ect were (1) subclinical, which are physiological effects only, such as the inhibition of δ -aminolevulinic acid dehydratase (ALAD; see Section AX7.1.3.3); (2) t oxic, a threshold level marki ng the initiation of clinical signs, such as anemia, lesions in tissues, weight loss, muscular incoor dination, green diarrhea, and anorexia; and (3) compatible with death, an approximate threshold value associated withAX7-43 death in field, captive, and/or experimental cases of Pb poisoning. The tissue Pb levels associated with these levels of effects are presented in Table AX7-1.2.1.

Table AX7-1.2.1. Tissue Lead Levels in Birds Causing Effects (taken from Franson, 1996)

Order Blood (μg/dL) Liver (ppm wet wt.) Kidney (ppm wet wt.) Falconiformes Subclinical 0.2 – 1.5 2 – 4 2 – 5
Toxic >1 >3 >3 Compatible with death >5 >5 >5 Columbiformes
Subclinical 0.2 – 2.5 2 – 6 2 – 20
Toxic >2 >6 >15 Compatible with death >10 >20 >40 Galliformes
Subclinical 0.2 – 3 2 – 6 2 – 20
Toxic >5 >6 >15 Compatible with death >10 >15 >50

Tissue residue levels below the subclini cal levels in Table AX7-1.2.1 should be considered “background” (Franson, 1996). Levels in the subclinical rang e are indicative of potential injury from which th e bird would probably recover if Pb exposure was terminated.

Toxic residues could lead to death. Residues above the compa tible-with-death threshold are consistent with Pb-poisoning mortality (Fra nson, 1996). Additional information on residue levels for Passeriformes (e.g., sparrows, starlings , robins, cowbirds), Charadriiformes (e.g., gulls, terns), Gruiformes (e.g., cranes), Ciconiformes (e.g., egrets), Gaviformes (e.g., loons), and Strigiformes (e.g., owls) is available in Fr anson (1996). Scheuhammer (1989) found blood Pb concentrations of between 0.18 and 0.65 μg/m L in mallards corresponded to conditions associated with greater than normal exposure to Pb but that should not be considered Pb poisoning.AX7-44 (AX7-3) Lead concentrations in various tissues of mammals also have b een correlated with toxicity (Ma, 1996). The tissues commonly analysed fo r Pb are blood, liver, and kidney. Typical baseline levels of blood Pb are approximately 4 to 8 μg/dL for small mammals, and 2 to 6 μg/dL for mature cattle. Typical baseline levels of Pb in liver are 1 to 2 mg/kg dw for small mammals. Typical baseline levels of Pb in kidney are 0.2 to 1.5 mg/kg dw for mice and voles, but shrews typically have higher baseline leve ls of 3 to 19 mg/kg dw. Ma ( 1996) concluded that Pb levels less than 5 mg/kg dw in liver a nd 10 mg/kg dw in kidney were not associated with toxicity, but that levels greater than 5 mg/kg dw in liver a nd greater than 15 mg/kg dw in kidney could be taken as a chemical biomarker of toxic exposur e to Pb in mammals. Humphreys (1991) noted

that the concentrations of Pb in liver and kidne y can be elevated in animals with normal blood Pb concentrations (and without exhibiting clinical signs of Pb toxicity), because Pb persists in these organs longer than in blood. Uptake of Pb from soil into small mammals was also modeled as part of Eco-SSL development (U.S. Environmental Protection Agen cy, 2005a). The relationship derived between Pb in the soil and Pb in the whole-body of a small mammal was taken from Sample et al. (1998)

and is as follows:

Ln(Cmammal) = 0.4422 * Ln(Csoil) + 0.0761

Where Cmammal is the concentration of Pb in small mammals (dry weight) and Csoil is the

concentration of Pb in the soil. Similar to the uptake equations fo r plants (Eq. 8-1) and earthworms (Eq 8-2), the equation for mammalia n uptake recognizes that the ratio of Pb

concentration in small mammals to Pb concentration in soil is not constant.

AX7.1.2.4 Resistance Mechanisms Many mechanisms related to heavy metal tolera nce in plants and invertebrates have been described, including avoidance (i.e., root redistribution, food rejecti on), exclusion (i.e., selective uptake and translocation), immobilization at th e plant cell wall, and excretion (i.e., foliar leakage, moulting) (Tyler et al., 1989; Patra et al., 2004). The following section reviews the

recent literature on the resistance mechanisms of plants and invertebrates through mitigation of Pb (1) toxicity or (2) exposure.AX7-45 Detoxification Mechanisms Lead sequestration in cell walls may be the most important detoxification mechanism in plants. Calcium may play a role in this detoxi fication by regulating intern al Pb concentrations through the formation of Pb-cont aining precipitates in the cell wall (Antosiewicz, 2005). Yang et al. (2000) screened 229 varieties of rice ( Oryza sativa ) for tolerance or sensitivity to Pb and

found that the oxalate conten t in the root and root exudates was in creased in Pb-tolerant varieties. The authors suggested that the ox alate reduced Pb bioavailability, and that this was an important tolerance mechanism (Yang et al., 2000). Sharma et al. (2004) found Pb-sul fur and Pb-sulfate in the leaves, and Pb-sulfur in the roots of Sesbania drummondii (Rattlebox Drummond), a Pb hyperaccumulator plant grown in Pb-nitrate solu tion. They hypothesized that these sulfur ligands were indicative of glutathione and phytoc helatins, which play a role in heavy metal

homeostasis and detoxificati on (Sharma et al., 2004). Sea pinks ( Armeria maritima ) grown on a metal-contaminated site (calamine spoils more than 100 years old) accumulated 6 Η the concentrations of Pb in brown (dead and withering) leaves than green leaves (Szarek -Lukaszewska et al., 2004). The concentration of Pb in brown leaves was similar to that in roots. This grea ter accumulation of Pb in to older leaves was not observed in plants grown hydroponically in the la boratory. The authors hypothesized that this sequestering of Pb into the oldest leaves wa s a detoxification mechan ism (Szarek-Lukaszewska et al., 2004). Terrestrial invertebrates also mitigate Pb t oxicity. Wilczek et al. (2004) studied two species of spider , the web-building Agelena labyrinthica, and the active hunter wolf spider Pardosa lugubris . The activity of metal detoxifying en zymes (via the glutathione metabolism pathways) was greater in A. labyrinthica and in females of both species (Wilczek et al., 2004). Marinussen et al. (1997) found th at earthworms can excrete 60% of accumulated Pb very quickly once exposure to Pb-contaminated soils has ended. However, the remainder of the body burden is not excreted, possibly due to the storage of Pb in waste nodules that are too large to be excreted (Hopkin, 1989). Gintenreit er et al. (1993) found that Le pidoptera larvae (in this case, the gypsy moth Lymantria dispar ) eliminated Pb, to some extent, in the meconium (the fluid excreted shortly after emergence from the chrysalis). Lead, in the form of pyromorphite (Pb 5 (PO 4 ) 3 Cl), was localized in the anterior pharynx region of the nematode Ceanorhabditis elegans (Jackson et al., 2005). The authors hypothesizedAX7-46 that the nematode may detoxify Pb via its precipitation into pyromorphite, which is relatively insoluble (Jackson et al., 2005).

Avoidance Response Studies with soil invertebrates hypothesize th at these organisms ma y avoid soil with high Pb concentrations. For example, Bengtsson et al. (1986) suggested that the lower Pb concentrations in earthworm tissues may be a result of lowered feeding activity of worms at higher Pb concentrations in soil.

AX7.1.2.5 Physiological Effects of Lead

Several studies have measured decrease d blood ALAD activity in birds and mammals exposed to Pb (U.S. Environmental Protec tion Agency, 1986a). Recent studies on the physiological effects of Pb to consumers have focused on heme synthesis (as measured by ALAD activity and protoporphyrin concentration) , lipid peroxidation, an d production of fatty acids. Effects on growth are covered in Section AX7.1.4. Biochemically, Pb adversely affects hemoglobi n synthesis in birds and mammals. Early indicators of Pb exposure in birds and mamm als include decreased blood ALAD concentrations and increased protoporphyrin IX activity. The ef fects of Pb on blood parameters and the use of these parameters as sensitive biomarkers of exposure has been well documented (Eisler, 1988; U.S. Environmental Protection Agency, 2005b). However, the linkage between these

biochemical indicators and ecol ogically relevant effects is less well understood. Low-level inhibition of ALAD is not generally considered a toxic response, because this enzyme is thought to be present in excess concentrations; rather , it may simply indicate that the organism has recently been exposed to Pb (Henny et al., 1991). Schlick et al. (1983) studied ALAD inhibiti on in mouse bone marrow and erythrocytes. They estimated that an absorbed dose of between 50 and 100 μg Pb-acetate/kg body weight per day would result in long-term inhibition of ALAD. Beyer et al. (2000) related blood Pb to s ublethal effects in waterfowl along the

Coeur d = Alene River near a mining site in Idaho. The sublethal effects measured included, among others, red blood cell ALAD activity and prot oporphyrin levels in th e blood. As found in other studies, ALAD activity was th e most sensitive indicator of Pb exposure, decreasing to 3%AX7-47 of the reference value at a blood Pb con centration of 0.68 mg/kg ww (wet weight). Protoporphyrin concentrations showed a 4.2-fold increase at this same concentration. Henny et al. (1991) studied osprey along the Coeur d = Alene River. There were no observations of death, behavioral abnormalities, or reduced productivity re lated to Pb exposure, although inhibition of blood ALAD and increased pr otoporphyrin concentrations were measured in ospreys. Henny et al. (1991) hypothesized th at no impacts to osprey were observed, even though swan mortality was documented in the area because swans feed at a lower trophic level (i.e., Pb does not biomagnify, and thus is found at higher concentrations in lower trophic level organisms). Hoffman et al. (2000a) also studied the e ffects of Coeur d’Alene River sediment on waterfowl, focusing on mallard ducklings for 6 w eeks after hatching. The study revealed that a 90% reduction in ALAD activity and a greate r than 3-fold increase in protoporphyrin

concentration occurred when blood Pb reached a concentration of 1.41 mg/kg ww as a result of the ducklings being fed a diet composed of 12% sediment (3449 mg/kg Pb). Those ducklings fed a diet composed of 24% sediment were f ound to have a mean blood Pb concentration of 2.56 mg/kg ww and a greater than 6-fold increase in protoporphyrin concentration. Hoffman et al. (2000b) also studied Canada Geese ( Branta canadensis ) goslings in a similar fashion. The results revealed that, while blood Pb concen trations in goslings we re approximately half (0.68 mg/kg ww) of those found in ducklings under the same conditions (12% diet of 3449 mg/kg sediment Pb), goslings showed an increased sensitivity to Pb exposure. Goslings experienced a 90% reduction in ALAD activit y and a 4-fold increase in protoporphyrin concentration, similar to conditions found in th e ducklings, although blood Pb concentrations

were half those found in the duckl ings. More serious effects we re seen in the goslings when blood Pb reached 2.52 mg/kg, including d ecreased growth an d mortality. Redig et al. (1991) reported a hawk LOAEL (lowest observed adverse effect level) of 0.82 mg/kg-day for effects on heme biosynthetic pathways. Lead dosages as high as 1.64 to 6.55 mg/kg-day caused neither mortal ity nor clinical signs of toxicity. A dose of 6.55 mg/kg-day resulted in blood Pb levels of 1.58 μg/mL. Th ere were minimal changes in immune function (Redig et al., 1991). Repeated oral administration of Pb resulted in biochemical alterations in broiler chickens (Brar et al., 1997a,b). At a dose of 200 mg/kg-day Pb -acetate, there were significant increases inAX7-48 plasma levels of uric acid and cr eatinine and significant declines in the levels of total proteins, albumin, glucose, and cholesterol. Brar et al. (1997a) suggested that increased uric acid and

creatinine levels could be due to an accelerated rate of protein catabolism and/or kidney damage. They also suggested that the decline in plasma proteins and albumin levels may be caused by diarrhea and liver dysfunction due to the Pb e xposure. Brar et al. (1997b) also found that significant changes in plasma enzymes may be causing damage to other organs. Lead can cause an increase in tissue lip id peroxides and changes in glutathione concentrations, which may be related to peroxi dative damage of cell membranes (Mateo and Hoffman, 2001). There are species-specific differen ces in resistance to oxidative stress (lipid peroxidation), which may explain why Canada g eese are more sensitive to Pb poisoning than mallards (Mateo and Hoffman, 2001). Lead also cau sed an increase in the production of the fatty acid arachidonic acid, which has been associated with changes in bone formation and immune response (Mateo et al., 2003a). The effects observed by Mateo et al. (2003a,b) were associated with very high concentrations of Pb in the diet (1840 mg Pb/kg diet), much higher than would be found generally in the environment, and high en ough that birds decreased their food intake. Lead also induces lipid peroxi dation in plants. Rice plants exposed to a highly toxic level of Pb (1000 μM in nutrient solution) showed elevat ed levels of lipid peroxides, increased activity of superoxide dismutase, guaiacol peroxidase, ascorbate peroxidase, a nd glutatione reductase (Verma and Dubey, 2003). The elevated levels of these enzymes suggest the plants may have an antioxidative defense mechanism against oxida tive injury caused by Pb (Verma and Dubey, 2003).

AX7.1.2.6 Factors that Modify Organism Response

Research has demonstrated that Pb ma y affect survival, reproduction, growth, metabolism, and development in a wide range of species. These effects may be modified by chemical, biological, and physical factors. The f actors that modify respons es of organisms to Pb are described in the following sections.

Genetics Uptake and toxicity of Pb to plants are influen ced strongly by the type of plant. Liu et al. (2003) found that Pb uptake and translocation by rice plants differed by cultivar (a cultivatedAX7-49 variety of plant produced by se lective breeding) but was not related to genotype. Twenty cultivars were tested from three genotypes. The differences in Pb concentrations among cultivars were smallest when comparing concentr ations in the grains at the ripening stage. This study also found that toxicity varied by cu ltivar; at 800 mg Pb/kg soil, some cultivars were greatly inhibited, some were significantly improved, and others showed no change. Dearth et al. (2004) compared the response of Fisher 344 (F344) rats and Sprague-Dawley (SD) rats to exposure via gavage to 12 mg Pb/mL as Pb-acetate. Blood Pb levels in the F344 dams were higher than those of the SD da ms. Lead delayed the timing of puberty and suppressed hormone levels in F344 offspring. Th ese effects were not obs erved in the offspring of SD rats, even when the dose was doubled. The authors conclude that F344 rats are more sensitive to Pb (Dear th et al., 2004).

Biological Factors Several biological factors may influence Pb uptake and organism response, including organism age, sex, species, feeding guild, and, fo r plants, the presence of mycorrhizal fungi. Monogastric animals are more sensitive to Pb than ruminants (Humphreys, 1991). Younger organisms may be more susceptible to Pb toxicity (Eisler, 1988; Humphreys, 1991). Nestlings are more sensitive to the effects of Pb than older birds, and young altricial birds (species unable to self-regulate body heat at birt h, such as songbirds) , are considered more sensitive than precocial birds (species that have a high degree of independe nce at birth, such as quail, ducks, and poultry) (Scheuhammer, 1991).

Gender can also have an effect on the accu mulation of Pb by wild life (Eisler, 1988). Female birds accumulate more Pb than males (Scheuhammer, 1987; Tejedor and Gonzalez, 1992). These and other authors ha ve related this to the increas ed requirement for calcium in laying females. Different types of invertebrates accumulate di fferent amounts of Pb from the environment (U.S. Environmental Protection Agency, 1986a). There may be species- and sex-specific

differences in accumulation of Pb into invertebra tes, specifically arthropods. This has been shown by Wilczek et al. (2004) who studied two species of spider, the web-building A. labyrinthica and the active hunter wolf spider P. lugubris . The body burdens of Pb in the wolf spider were higher than in the web-build ing spider, and this ma y be due to the moreAX7-50 effective use of glutathione metabolism pathways in A. labyrinthica . Body burdens of females

were lower than those of males in both species. This was also observed in spiders by Rabitsch (1995a). Females are thought to be able to deto xify and excrete excess metals more effectively than males (Wilczek et al., 2004). Lead accumulation has been measured in numerous species of arthropods with different feed ing strategies. Differences were observed between species (Janssen and Hogervorst, 1993; Rabitsch, 1995a) and depending upon sex (Rabitsch, 1995a), developmental stage (Gintenreiter et al., 1993; Rabitsch, 1995a), and seas on (Rabitsch, 1995a). Uptake of Pb may be enhanced by symbio tic associations between plant roots and mycorrhizal fungi. Similar to the mechanism a ssociated with increased uptake of nutrients, mycorrhizal fungi also may cause an increase in the uptake of Pb by increasing the surface area of the roots, the ability of the root to absorb pa rticular ions, and the tran sfer of ions through the soil (U.S. Environmental Protection Agency, 1986a). There have been contradictory results published in the literature rega rding the influence of mycorrhizal organisms on the uptake and toxicity of Pb to plants (see review in Påhlsson, 1989). Lin et al . (2004) found that the bioavailability of Pb increased in the rhizosphere of rice plants, although th e availability varied with Pb concentration in soil. Bioavailability wa s measured as the soluble plus exchangeable Pb fraction from sequential extrac tion analysis. The authors hyp othesized that the enhanced solubility of Pb may be due to a reduced pH in the rhizosphere or, more likely, the greater availability of organic ligands, which further s timulates microbial growth (Lin et al., 2004). Increased bioavailability of Pb in soil may increase the uptak e of Pb into plants, although this was not assessed by Lin et al. (2004). However, Di xon (1988) found that red oak ( Quercus rubra ) seedlings with abundant ectomycorrhizae had lower Pb concentrations in their roots than those seedlings w ithout this fungus, although only at the 100 mg Pb/kg sandy loam soil concentration (no differences were found at lowe r Pb concentrations). Lead in soil also was found to be toxic to the ectomycorrhizal fungi after 16 weeks of exposure to 50 mg Pb/kg or more (Dixon, 1988). Malcova and Gryndler (2003) showed that maize root exudates from mycorrhizal fungi can ameliorate heavy metal toxi city until a threshold metal concentration was surpassed. This may explain the conflicting results in the past regarding the uptake and toxicity of Pb to plants with mycorrhizal fungi. The type of food eaten is a major determinan t of Pb body burdens in small mammals, with insectivorous animals accumulating more Pb than herbivores or granivores (U.S. EnvironmentalAX7-51 Protection Agency, 1986a). In fact, the main i ssue identified by the EPA (U.S. Environmental Protection Agency, 1986a) related to invertebra te uptake of Pb was not toxicity to the invertebrates, but accumulation of Pb to levels th at may be toxic to their consumers. Several authors suggest that shrews are a good indicato r of metal contamination, because they tend to accumulate higher levels of metals than herb ivorous small mammals (see data summary in Sample et al. (1998)). Shrews accumulate higher levels of metals in contaminated habitats, because their diet mainly consists of detritivores (i.e., earthworms) and other soil invertebrates in

direct contact with the so il (Beyer et al., 1985).

Physical/Environmental Factors

Plants The uptake and distribution of Pb into higher plants from the soil is affected by various chemical and physical factors including the chemi cal form of Pb, the pr esence of other metal ions, soil type, soil pH, cati on exchange capacity (CEC), the amount of Fe/Mn-oxide films present, organic matter content, temperature, ligh t, and nutrient availability. A small fraction of Pb in soil may be released into the soil water, which is then available to be taken up by plants (U.S. Environmental Protection Agency, 1986a).

The form of Pb has an influen ce on its toxicity to plants. Fo r example, Pb-oxide is less

toxic than more bioavailable forms such as Pb -chloride or Pb-acetate. In a study by Khan and Frankland (1983), radish plants were exposed to Pb-oxide and Pb-chloride in a loamy sand at pH 5.4, in a 42-day study. In a tested concentrati on range of 0 to 5000 mg/kg, root growth was inhibited by 24% at 500 mg/kg for Pb-chloride and an EC 50 of 2400 mg/kg was calculated from a dose-response curve. Plant growth ceased at 5000 mg/kg and shoots exhibited an EC 50 of 2800 mg/kg. For Pb-oxide exposure (concentration range of 0 to 10,000 mg/kg), reported results

indicate an EC 50 of 12,000 mg/kg for shoot growth and an EC 50 of 10,000 mg/kg for root growth. There was no effect on root growth at 500 mg/ kg and a 26% reduction at 1000 mg/kg Pb oxide. Soil pH is the most influential soil property wi th respect to uptake and accumulation of Pb into plant species. This is most likely due to increased bioavailability of Pb created by low soil pH. At low soil pH conditions, markedly elevat ed Pb toxicity was reported for red spruce ( P. rubens ) (Seiler and Paganelli, 1987). At a soil pH of 4.5, ryegrass ( Lolium hybridum )AX7-52 and oats ( Avena sativa ) had significantly higher Pb concentr ations after 3 months of growth compared to plants grown at pH 6.4 (Allinson and Dzialo, 1981).

Invertebrates The uptake of Pb into invert ebrates depends on the physical environment and parameters such as pH, calcium concentration, organic matte r content, and CEC. Greater accumulation is found generally when the soil pH or organic content is lower (U.S. Environmental Protection Agency, 1986a). Soil pH has a significant influence on uptake of Pb into invertebrates. Perämäki et al. (1992) studied the influence of so il pH on uptake into the earthworm Aporrectodea caliginosa

. Lead accumulation was lowest at the highest pH values, but there was no statistical difference due to variability in the data. Variability in the response also was found by Bengtsson et al. (1986), who reared earthworms ( Dendrobaena rubida ) in acidified soils at pH 4.5, 5.5, or 6.5. Lead uptake into worms was pH-dependent, although the highest concentrations were not always found at the lowest pH. There was no clear re lationship between Pb concentration in cocoons and soil pH, and Pb concentrations were higher in the hatchlings than in the cocoons. As has been reported in many other studies (Neuhauser et al., 1995), concentration fa ctors (ratio of Pb in worm to Pb in soil) were lower at higher Pb conc entrations in soil. The authors attribute some of this to a lowered feeding activity in worms at hi gher Pb concentrations (Bengtsson et al., 1986). Beyer et al. (1987) and Morgan and Morgan (1 988) recognized that other factors beyond soil pH could influence the uptake of Pb into earthworms, which may be the cause of the inconsistencies reported by several authors. Both studies ev aluated worm uptake of Pb relative to pH, soil calcium concentration, and organic ma tter content. Morgan and Morgan (1988) also considered CEC, and Beyer et al. (1987) consid ered concentrations of phosphorus, potassium, or magnesium in soil. Both studies found that calc ium concentrations in soil were correlated with soil pH. Morgan and Morgan (1988) also f ound that CEC was correlated with percentage organic matter. Soil pH (coupled with CEC) a nd soil calcium were found to play significant roles in the uptake of Pb into worms (Beyer et al., 1987; Morgan and Morgan, 1988). Beyer et al. (1987) noted that concentratio ns of phosphorus in soil had no effect.AX7-53 Nutritional Factors Diet is a significant modifier of Pb absorp tion and of toxic effect s in many species of birds and mammals (Eisler, 1988). Dietary deficiencies in calcium, zinc, iron, vitamin E, copper, thiamin, phosphorus, magnesium, fat, protein, mi nerals, and ascorbic acid increased Pb absorption and its toxic effects (Eisler, 1988). Mateo et al. (2003b) studied in traspecies sensitivity to Pb-induced oxidative stress, by varying the vitamin E content of mallard diets. Vitamin E can protect against peroxidative damage and was found to decrease the lipid peroxidation in nerves of birds; how ever, it did not alleviate any sign of the Pb poisoning. The au thors hypothesize that inhibition of antioxidant enzymes and interaction with sulfhydryl groups of proteins may have a greater influence on Pb toxicity than lipid peroxidati on (Mateo et al., 2003b). The e ffects observed by Mateo et al. (2003b) were associated with very high concentrati ons of Pb in diet (1840 mg Pb/kg diet), much higher than would be found gene rally in the environment, a nd high enough that the birds

decreased their food intake. Mallard ducklings were exposed to Pb-contam inated sediment and either a low nutrition or optimal nutrition diet (Dougl as-Stroebel et al., 2005). Le ad exposure combined with a nutritionally inferior diet caused more changes in behavior (as measured by time bathing, resting, and feeding) than Pb exposure or low-nutrition di et alone. These effects may be due to the lownutrition diet being deficient in le vels of protein, amino acids, calci um, zinc, and other nutrients. Zebra finches ( Taeniopygia guttata ) were exposed to Pb-acetate via drinking water at 20 mg/L for 38 days, along with ei ther a low- or high-calcium diet (Snoeijs et al., 2005). Lead uptake into tissues was enhanced by a low-calc ium diet. Lead did not affect body weight, hematocrit, or adrenal stress response. Lead suppressed the humoral immune response only in females on a low-calcium diet, suggesting increase d susceptibility of females to Pb (Snoeijs et al., 2005).

Interactions with Other Pollutants Lead can interact with other pollutants to ex ert toxicity in an an tagonistic (less than additive), independent, additive, or synergistic (more than additive) manner. Concurrent exposure to Pb and additional pollu tant(s) can affect the ability of plants to uptake Pb or the other pollutant. However, the up take and toxic response of plants exposed to Pb combined withAX7-54 other metals is inconsistent (P åhlsson, 1989). Therefore, no gene ralizations can be made about the relative toxicity of metal mixtures. For example, An et al. (2004) conducted acute, 5-day bioassays on cucumber exposed to Pb, Pb + c opper, Pb + cadmium, or Pb + copper + cadmium in a sandy loam soil of pH 4.3. Shoot and root growth were measured. Depending on the tissue and metal combination, additivity, synergism, or antagonism was observed in the responses to these metals. In fact, the response in roots was no t consistent with the re sponse in shoots for the binary mixtures. However, the combined effects were greater in the roots than the shoots, which may be explained by the tendency for Pb and other heavy metals to be retained in the roots of plants. In addition, the pattern of metal bioaccumulation into plant tissue did not always correlate with the toxic response. However, an tagonism was observed in the response of roots and shoots exposed to all three metals, and this was reflected in the decreased accumulation of metals into plant tissues. The authors hypothesized that this may be due to the formation of less bioavailable metal complexes (An et al., 2004). He et al. (2004) found that selenium and zinc both inhibited the uptake of Pb into Chinese cabbage ( Brassica rapa ) and lettuce ( Lactuca sativa ). Zinc applied at 100 mg/kg or selenium applied at 1 mg/kg decreased the uptake of Pb (p resent in soil at 10 mg /kg as Pb-nitrate) into lettuce by 15% and 20%, respectively, and into Chinese cabbage by 23 and 20%, respectively. Selenium compounds were evaluated to de termine whether they could change the inhibition of ALAD in liver, kidne y, or brain of mice exposed to Pb-acetate (Perottoni et al., 2005). Selenium did not affect the inhibition of ALAD in the kidney or liver, but it did reverse

the ALAD inhibition in mouse brain.

Co-occurrence of cadmium with Pb resulted in reduced blood Pb concentrations in rats (Garcia and Corredor, 2004). The authors hypothe sized that cadmium ma y block or antagonize the intestinal absorption of Pb, or the metallo thionein induced by cadmium may sequester Pb. However, this was not observed in pigs, wher e blood Pb concentrations were greater when cadmium was also administered (Phillips et al., 2003). The effect on growth rate also was additive when both metals were given to young pigs (Phillip s et al., 2003).AX7-55 AX7.1.2.7 Summary The current document expands upon and upda tes knowledge related to the uptake, detoxification, physiological effect s, and modifying factors of Pb toxicity to terrestrial organisms.

Surface Deposition onto Plants Recent work (Dalenberg and Van Driel, 1990; Jones and Johnston, 1991; Angelova et al., 2004) has supported previous results and conclu sions that surface deposition of Pb onto aboveground vegetation from airborne sources may be significant (U.S. Environmental Protection Agency, 1986a). Similarly, it has been well documente d previously that Pb in soil also is taken up by plants, although most remains in the roots, ther e is little translocatio n to shoots, leaves, or other plant parts (U.S. Environmental Protecti on Agency, 1986a). More recent work continues to support this finding (Sieghardt, 1990), and one study found increase d tolerance in species with bulbs, possibly due to the storage of Pb in the bulb (Wierzbicka, 1999).

Uptake Mechanism into Plants Lead was thought previously to be taken up by plants via th e symplastic route (through cell membranes), although it was unknown whethe r some Pb also may be taken up via the apoplastic route (between cells ) (U.S. Environmental Protecti on Agency, 1986a). Recent work has shown that the apoplastic ro ute of transport is stopped in th e primary roots by the endodermis (Sieghardt, 1990), supporting the pr evious conclusion that the sy mplastic route is the most significant route of tran sport into plant cells.

Species Differences in Uptake into Earthworms Different species of earthworm accumulated di fferent amounts of Pb, and this was not related to feeding strategy (U.S . Environmental Protection Agenc y, 1986a). This is supported by recent work, which has shown Aporrectodea accumulated more than Lumbricus (Terhivuo et al., 1994; Pižl and Josens, 1995), al though this is not consistent ly observed (Spurgeon and Hopkin, 1996a).AX7-56 Speciation and Form of Lead Recent work supports previous conclusions that the form of metal tested, and its speciation in soil, influence uptak e and toxicity to plants and in vertebrates (U.S. Environmental Protection Agency, 1986a). The oxide form is less toxic that the chlori de or acetate forms,

which are less toxic than the ni trate form of Pb (Khan and Frankland, 1983; Lock and Janssen, 2002; Bongers et al., 2004). Howeve r, these results must be inte rpreted with caution, as the counterion (e.g., the nitrate ion) may be contribut ing to the observed to xicity (Bongers et al., 2004).

Detoxification in Plants Lead may be deposited in root cell wa lls as a detoxificat ion mechanism (U.S. Environmental Protection Agency, 1986a), and this may be influenced by calcium concentrations (Antosiewicz, 2005). Yang et al. (2000) suggested that the oxalate content in root and root exudates reduced the bioavailability of Pb in so il, and that this was an important tolerance mechanism. Other hypotheses put forward recen tly include the presence of sulfur ligands (Sharma et al., 2004) and the sequestration of Pb in old leaves (Szarek-L ukaszewska et al., 2004)

as detoxification mechanisms.

Detoxification in Invertebrates

Lead detoxification has not been studied ex tensively in invertebrates. Glutathione detoxification enzymes were measured in two speci es of spider (Wilczek et al., 2004). Lead may be stored in waste nodules in earthworms (H opkin, 1989) or as pyromorphite in the nematode (Jackson et al., 2005).

Physiological Effects The effects on heme synthesis (as meas ured by ALAD activity and protoporphyrin concentration, primarily) have been well-docum ented (U.S. Environmental Protection Agency, 1986a) and continue to be studied (Schlick et al., 1983; Scheuhammer, 1989; Henny et al., 1991; Redig et al., 1991; Beyer et al ., 2000; Hoffman et al., 2000a,b). However, Henny et al. (1991) caution that changes in ALAD and other enzyme parameters are not always related to adverse effects, but simply indicate exposure. Other effects on plasma enzymes, which may damageAX7-57 other organs, have been reported (Brar et al., 1997 a,b). Lead also may cause lipid peroxidation (Mateo and Hoffman, 2001), which may be alle viated by vitamin E, although Pb poisoning may still result (Mateo et al., 2003b). Changes in fatty acid production have been reported, which may influence immune response and bone formation (Mateo et al., 2003a).

Response Modification Genetics, biological factors, phys ical/environmental factors, nu tritional factors, and other pollutants can modify terrestrial organism response to Pb. Fisher 344 rats were found to be more sensitive to Pb than Sprague-Dawley rats (D earth et al., 2004). Y ounger animals are more sensitive than older animals (E isler, 1988; Scheuhammer, 1991), and females generally are more sensitive than males (Scheuhammer, 1987; Teje dor and Gonzalez, 1992; Snoeijs et al., 2005). Monogastric animals are more sensitive than ruminants (Humphreys, 1991). Insectivorous mammals may be more exposed to Pb than herb ivores (Beyer et al., 1985; Sample et al., 1998), and higher tropic-level consumers may be less exposed than lower trophic-level organisms (Henny et al., 1991). Diets defici ent in nutrients (including calc ium) result in increased uptake of Pb (Snoeijs et al., 2005) and greater toxicity (Douglas-Stroebel et al., 2005) in birds, relative to diets containing adequa te nutrient levels. Mycorrhizal fungi may ameliorate Pb toxicity until a threshol d is surpassed (Malcova and Gryndler, 2003), which may explain why some studies show increase d uptake into plants (Lin et al., 2004) while others show no difference or less uptake (Di xon, 1988). Uptake of Pb into plants and soil invertebrates increases with a de crease in soil pH. However, calcium content, organic matter content, and ca tion exchange capacity of soil s also have had a significant influence on uptake of Pb into plants and inve rtebrates (Beyer et al., 1987; Morgan and Morgan, 1988).

Interactions of Pb with other metals ar e inconsistent, depending on the endpoint measured, the tissue analyzed, the animal species , and the metal combination (Phillips et al., 2003; An et al., 2004; He et al ., 2004; Garcia and Corredor, 2004; Perottoni et al., 2005).



### AX7.1.3 Exposure-Response of Terrestrial Species Section AX7.1.3 summarized the most important factors related to uptake of Pb by terrestrial organisms, the physiological effects of Pb, and the factors th at modify terrestrialAX7-58 organism responses to Pb. Section AX7.1.4 ou tlines and highlights the critical recent advancements in the understanding of the toxicity of Pb to terrestrial organisms. This section begins with a summary of the conclusions from the 1986 Lead AQCD (U.S. Environmental Protection Agency, 1986a) and then summarizes th e more recent critical research conducted on effects of Pb on primary producer s, consumers, and decomposer s. All concentrations are expressed as mg Pb/kg soil dw, unless otherwise indicated. Lead exposure may adversely affect organisms at different levels of organization, i.e., individual organisms, populations, communities, or ecosystems. Generally, however, there is insufficient information available for single mate rials in controlled studies to permit evaluation of specific impacts on higher levels of organizatio n (beyond the individual organism). Potential effects at the population level or higher are, of necessity, extrapolated from individual level studies. Available population, community, or ecos ystem level studies are typically conducted at sites that have been contaminated or adversely affected by multiple stressors (several chemicals alone or combined with physical or biological stressors). Theref ore, the best documented links

between lead and effects on the environment are w ith effects on individual organisms. Impacts on terrestrial ecosystems are discusse d in Section 7.1.5 and Annex AX7.1.5. The summary of recent critical advancements in understanding toxicity relies heavily on the work completed by a multi-stakeholder group, consisting of federal, state, consulting, industry, and academic participants, led by the EP A to develop Ecological Soil Screening Levels (Eco-SSLs). Eco-SSLs describe th e concentrations of contaminants in soils that would result in little or no measurable effect on ecological r eceptors (U.S. Environmen tal Protection Agency, 2005a). They were developed by the U.S. EPA for use in screening-level assessments at Superfund sites to identify contaminants requir ing further evaluation in an ecological risk assessment and were not designed to be used as cleanup target levels. The Eco-SSLs are intentionally conservative in orde r to provide confidence that c ontaminants, which could present an unacceptable risk, are not screened out early in the evaluation process. That is, at or below these levels, adverse effects are considered unl ikely. Eco-SSLs were derived for terrestrial plants, soil invertebrates, birds, and mammals. Detailed procedur es using an ex tensive list of acceptability and exclusion criteria (U.S. Enviro nmental Protection Agency, 2005a) were used in screening the toxicity studies to ensure that only those that met minimum quality standards were used to develop the Eco-SSLs. In addition, two peer reviews were completed during theAX7-59 Eco-SSL development process. The first was a consultation with the EPA Science Advisory Board (SAB) in April 1999, and the second was a peer review workshop in July 2000, which was open to the public.

Several conservative factors we re incorporated into the de velopment of the Eco-SSLs. For the plant and invertebrate Eco-SSLs, st udies were scored to favor relatively high bioavailability. For wildlife Eco-SSLs, only sp ecies with a clear exposure link to soil were considered (generalist species, species with a link to the aquatic environment, or species which consume aerial insects were excluded), simple diet classifications were used (100% plants, 100% earthworms or 100% animal prey) when in reality wildlife consume a vari ed diet, species were assumed to forage exclusively at the contaminated site, relative bioavailab ility or Pb in soil and diet was assumed to be 1, and the TRV was sel ected as the geometric mean of NOAELs unless this value was higher than the lowest bounded LOAEL for mortality, growth or reproduction (U.S. Environmental Protection Agency, 2005a,b).

Areas of research that were not addressed are effects from irrelevant exposure conditions relative to airborne emissions of Pb (e.g., Pb shot , Pb paint, injection st udies, studies conducted on mine tailings, and studies conducted with hydr oponic solutions); mixture toxicity (addressed in Section AX7.1.3); issues related to indirect ef fects (e.g., effects on pred ator/prey interactions, habitat alteration, etc.); and human health-related research (e.g., hypertension), which is

addressed in othe r sections of this document. The toxicity data presented herein should be reviewed w ith a note of caution regarding

their relevance to field conditions. Laboratory studies, partic ularly those using Pb-spiked soil, generally do not allow the soil to equilibrate following the addition of Pb and prior to the addition of test organisms. This may result in increased bioavailability and overestimation of Pb toxicity relative to actual environmen tal conditions (Davies et al., 2003).

AX7.1.3.1 Summary of Conclusions from the 1986 Lead Criteria Document The 1996 Lead AQCD, Volume II (U.S. Environmental Protection Agency, 1986a) reviewed the literature on the toxicity of Pb to plants, soil organisms, birds, and mammals. The main conclusions from that document are provided below.AX7-60 Primary Producers Commonly reported effects of Pb on vascul ar plants include the inhibition of photosynthesis, respiration, and/or cell elongation, all of which reduce plant growth. However, it

was noted that studies of other effects on plant processes such as maintenance, flowering, and hormone development had not been conducted; therefore, no conclusion could be reached concerning effects of Pb on these processes. The EPA (U.S. Environmental Protection Ag ency, 1986a) concluded that most plants experience reduced growth when Pb concentrati ons in soil moisture (the film of moisture surrounding soil particles in the root zone of soil) exceed 2 to 10 mg/kg. It also was concluded that most plants would experien ce reduced growth (inhibition of photosynthesis, respiration, or cell elongation) in soils of ∃ 10,000 mg/kg when soil composition and pH are such that bioavailability of Pb in the soil is low (see Section AX7.1.3 for details on factors affecting bioavailability of Pb in soil). Acid soils or soils with low organic matter tend to increase Pb bioavailability and would inhibit plants at mu ch lower Pb concentrations (e.g., as low as <100 mg/kg). Many effect levels have been reported at Pb concentrations much lower than 10,000 mg/kg soil. For example, effects on rye grass ( Lolium rigidum ) exposed to Pb in soil included inhibition of germinating root elongatio n (at <2.5 mg/kg), absence of root growth (at 5 mg/kg), or 55% inhibition of seed germin ation (at 20 to 40 mg/kg) . Stunted growth in radish ( Raphanus sativus ) was observed at 1000 mg/kg soil, with complete growth inhibition at 5000 mg/kg, when Pb was added as Pb-chloride; e ffects were less severe when the Pb was added

as Pb-oxide.

Consumers The EPA (U.S. Environmental Protection Ag ency, 1986a) concluded that food is the largest contributor of Pb to animals, with inha lation rarely accounting for more than 10 to 15% of daily intake of Pb and drinking water exposures being quite low. It al so was concluded that a regular dose of 2 to 8 mg/kg-da y causes death in most animals. Grazing animals may consume more than 1 mg/kg-day in habitats near sme lters and roadsides, but no toxic effects were

documented in these animals.AX7-61 Decomposers Lack of decomposition has been observed as a particular problem around smelter sites. Lead concentrations between 10,000 and 40,000 mg/kg soil can eliminate populations of decomposer bacteria and fungi (U.S. Envir onmental Protection Agen cy, 1986a). Lead may affect decomposition processes by direct toxici ty to specific groups of decomposers, by deactivating enzymes excreted by decomposers to break down organic matter, or by binding with the organic matter and rendering it resistan t to the action of decomposers. Microorganisms are more sensitive than plants to Pb in soil. Delayed decomposition may occur at between 750 and 7500 mg/kg soil (depe nding on soil type and other conditions).

Nitrification is inhibited by 14% at 1000 mg/kg soil.

U.S. Environmental Protection Agency Staff Review of 1986 Criteria Document The EPA reviewed the 1986 Lead AQCD a nd presented an overall summary of conclusions and recommendations (U.S. Enviro nmental Protection Agency, 1990). The major

conclusion was that available labora tory and field data indicated th at high concentrations of Pb can affect certain plants and alter the compositi on of soil microbial communities. It was noted

that few field studies were available in which Pb exposures and associated effects in wildlife were reported.

AX7.1.3.2 Recent Studies on the Effect s of Lead on Primary Producers Several studies published since 1986 have reported terrestrial pl ant exposure to Pb in soil, many of which were reviewed during the deve lopment of the Eco-SSLs (U.S. Environmental Protection Agency, 2005b). The relevant info rmation from the Eco-SSL document (U.S. Environmental Protection Agency, 2005b) is summari zed below. A literature search and review also was conducted to identify critical papers published since 2002, which is when the literature search was completed for Eco-SSL development, and no new papers were identified as critical to the understanding of Pb toxicity to terrestrial primary producers.

Effects observed in studies conducted since the 1986 Lead AQCD are similar to those reported previously and include decreased photos ynthetic and transpirati on rates and decreased growth and yield (U.S. Envir onmental Protection Agency, 2005b). The phytotoxicity of Pb is considered relatively low, due to the limited availability and uptake of Pb from soil and soilAX7-62 solution and minimal translocation of Pb from roots to shoots (Påh lsson, 1989). Although many laboratory toxicity studies have reported effects on plants, there are few reports of phytotoxicity from Pb exposure under field conditions. For ex ample, Leita et al. (1989) and Sieghardt (1990) reported high concentrations of Pb and othe r metals in soil and ve getation collected around mining areas in Europe, with no toxicity sy mptoms observed in plants or fruit. The literature search complete d for the terrestrial plant Eco-SSL development identified 439 papers for detailed review, of which 28 me t the minimum criteria (U.S. Environmental Protection Agency, 2005a). Thirty ecotoxicol ogical endpoints were gleaned from these 28 papers and were further evalua ted; most of those evaluate d growth (biomass), which was considered the most sensitive and ecologically re levant endpoint (U.S. Environmental Protection Agency, 2005b). Five of the endpoints, representi ng four species tested under three different combinations of pH and organic matter content, were used to develop the Eco-SSL of 120 mg/kg (115 mg/kg rounded to two significa nt digits) (Table AX7-1.3.1).

Table AX7-1.3.1. Plant Toxicity Da ta Used to Develop the Eco-SSL

Plant Species Soil pH % Organic Matter Toxicity Parameter Pb in Soil (mg/kg dw) Loblolly pine ( Pinus taeda ) 4 2.5 MATC * (growth) 144 Red maple ( Acer rubrum ) 4 2.5 MATC (growth) 144 Berseem clover ( Trifolium alexandrium ) 6.3 0.94 MATC (growth) 316

Berseem clover 6.7 3.11 MATC (growth) 141 Rye grass ( Lolium rigidum ) 5.6 0.1 MATC (growth) 22 Geometric Mean 115

*MATC = Maximum Acceptable Threshold Concentrat ion, or the geometric mean of the NOAEC (no-observed-adverse-effect concentration) and LOAE C (lowest-observed-adverse-effect concentration).

Source: U.S. Environmental Protection Agency (2005b).

The 25 ecotoxicological endpoints that were no t used to develop the Eco-SSL for plants are presented in Table AX7-1.3.2. The first six e ndpoints were considered eligible for Eco-SSL derivation but were not used; the remainder did not meet all of the requirements to be considered for inclusion in the Eco-SSL derivation process.AX7-63 Table AX7-1.3.2. Plant Toxicity Data Not Used to Develop the Eco-SSL

Plant Species Soil pH % Organic Matter Toxicity Parameter Pb in Soil (mg/kg dw) Studies eligible for Eco-SSL derivation, but not used

Berseem clover ( Trifolium alexandrium ) 6.7 3.11 MATC 141

Tomato ( Lycopericon esculentum ) 7.73 1.70 MATC 71

Tomato 8.20 0.86 MATC 71 Fenugreek ( Trigonella foenum-graecum ) 8.3 0.5 MATC 283

Spinach ( Spinacea oleracea ) 6.7 3.0 MATC 424 Corn ( Zea mays ) 6.5 2.1 MATC 158 Sow thistle ( Sonchus oleraceus ) 7.23 1.6 MATC 2,263 Studies not eligible for Eco-SSL derivation Loblolly pine ( Pinus taeda ) 5.5 3.4 NOAEC 480 Red oak ( Quercus rubra ) 6 1.5 LOAEC 100 Spinach 6.7 0.0 NOAEC 600 Alfalfa ( Medicago sativa ) 6.4 1.0 NOAEC 250

Alfalfa 6.9 1.7 NOAEC 250

Alfalfa 6.9 1.7 NOAEC 250 Radish ( Raphanus sativus ) 6.9 1.0 LOAEC 500

Radish 6.9 1.0 LOAEC 100

Radish 6.9 1.0 LOAEC 100 Onion ( Allium cepa ) 8.3 0.5 LOAEC 50

Radish 5.1 8.0 NOAEC 600 Carrot ( Daucus carota ) 7.0 0.6 NOAEC 85 Peas ( Pisum sativum ) 7.0 0.6 NOAEC 85 Barley ( Hordeum vulgare ) 6.0 2.5 NOAEC 1,000

Alfalfa 6.9 4.8 NOAEC 250

Tomato 7.45 2.06 MATC 35 Spinach 6.7 8.0 NOAEC 600

Radish 6.2 8.0 NOAEC 600

Radish 7.1 8.0 NOAEC 600

*MATC = Maximum Acceptable Threshold Concentr ation, or the geometric mean of the NOAEC (no-observed-adverse-effect concentration) and L OAEC (lowest-observed-advers e-effect concentration).

Source: U.S. Environmenta l Protection Agency (2005b).AX7-64 AX7.1.3.3 Recent Studies on the E ffects of Lead on Consumers Since the 1986 Lead AQCD (U.S. Environmen tal Protection Agency, 1986a), there have

been several studies in which birds and mammals were exposed to Pb via ingestion (primarily through dietary Pb). Many of these were review ed during development of the Eco-SSLs (U.S. Environmental Protection Agency, 2005b). The relevant information from the Eco-SSL document (U.S. Environmental Protection Agenc y, 2005b) is described below. A literature search and review was conducted to identify cr itical papers publishe d since 2002. These recent critical papers are described briefly below. No studies were found that used inhalation exposures to evaluate endpoints such as survival, grow th, and reproduction in birds or mammals. All studies described below exposed organisms via i ngestion (drinking water or diet) or gavage.

The Eco-SSLs for avian and mammalian consumers are presented as Pb concentrations in soil. These concentrations were calculated by assuming exposure to Pb via incidental soil ingestion and ingestion of Pb-contaminated food, and using a NOAEL as the TRV (U.S. Environmental Protection Agency, 2005a). A s implified version of the equation used to

calculate the Eco-SSL is: 

# ( ) 

# ( ) 

# [ ]

TRV BW IR C IR C HQ food food soil soil ÷ × + ×

=

where: HQ = hazard quotient (1 mg Pb/kg bw/day) C soil = concentration of Pb in soil (mg Pb/kg soil) IR soil = incidental soil ingestion rate (kg soil/day) C food = concentration of Pb in food (mg Pb/kg food) IR food = food ingestion rate (kg food/day) BW = body weight (kg) TRV = toxicity referen ce value (mg Pb/kg bw/day)

Food ingestion was estimated by modeling the uptake of Pb from soil into each diet component (e.g., vegetation, invertebrates, etc.). Bioavailability of Pb in soil and food was assumed to be 100%. The Eco-SSL is equivalent to the concentration of Pb in soil that results in an HQ = 1. The two factors that may have the most significant influe nce on the resulting EcoSSL are the assumption of 100% bioa vailability of Pb in soil and diet and the selection of the (AX7-4)AX7-65 TRV. The toxicity data that were reviewed to develop the TRV are presented in the following

subsections. Representative avian and mammalian wildlife species were selected for modeling Pb exposures to wildlife with di fferent diets and calculating the Eco-SSL. The avian species selected were dove (herbivore), woodcock (insec tivore), and hawk (carnivore). The mammalian species selected were vole (her bivore), shrew (insectivore), and weasel (carnivore). The lowest

of the three back-calculated soil concentrations, which resulted in an HQ = 1, was selected as the Eco-SSL. For Pb, the lowest values were for the insectivorous species of bird and mammal.

Avian Consumers

Effects on birds observed in studies conducted since the 1986 Lead AQCD (U.S. Environmental Protection Agency, 1986a) are simila r to those reported previously: mortality, changes in juvenile growth ra te and weight gain, effects on various reproductive measures, and changes in behavior (U.S. Environmental Pr otection Agency, 2005b). Reproductive effects following Pb exposure included declines in clut ch size, number of young hatched, and number of young fledged as well as decreased fertility or eg gshell thickness. Few significant reproductive effects have been reported in birds at Pb concentrations below 100 mg/kg in the diet (Scheuhammer, 1987). The literature search completed for Eco-SSL development identified 2,429 papers for detailed review for either avian or mammalian sp ecies, of which 54 met the minimum criteria for

further consideration for avian Eco-SSL develo pment (U.S. Environmental Protection Agency, 2005b). The 106 toxicological data points for bi rds that were further evaluated included biochemical, behavioral, physiological, pathological, reproductive, growth, and survival effects. Growth and reproduction data were used to derive the Eco-SSL (Table AX7-1.3.3; Figure AX7-1.3.1). The geometric mean of the NOAELs was calculated as 10.9 mg/kg-day, which was higher than the lowest bounded L OAEL (the term “bounded” means that both a

NOAEL and LOAEL were obtained from the same study). Therefore, the highest bounded

NOAEL that was lower than the lowest bounded LOAEL for survival, growth, or reproduction (1.63 mg Pb/kg bw-day) was used as the TRV (U.S. Environmental Protection Agency, 2005b).

The TRV was used to back-calculate the Eco- SSL of 11 mg/kg soil for avian species (U.S.AX7-66Table AX7-1.3.3. Avian Toxicity Data Used to Develop the Eco-SSLAvianSpeciesNo. ofDosesRoute ofExposureExposureDurationDurationUnitsAgeAgeUnitsLifestage SexEffectTypeEffectMeasureResponseSiteNOAEL(mg/kgbw/day)LOAEL(mg/kgbw/day)ReproductionJapanesequail4 FD 5 w 6 wLBF REP PROG WO 0.194 1.94Chicken 3 FD 4 w NR NR LB F REP PROG WO 1.63 3.26Chicken 4 FD 30 d 22 w LB F EGG ALWT EG 2.69 4.04Mallard 2 FD 76 d NR NR SM F EGG ESTH EG 5.63Americankestrel3 FD 6 mo 1-6 yr AD F REP RSUC WO 12.0Japanesequail5 FD 5 w 6 d JV M REP TEWT TE 12.6 126Japanesequail5 FD 5 w 1 d JV M REP TEWT TE 67.4 135Japanesequail3 FD 32 d NR NR AD F REP PROG WO 125Japanesequail5 FD 12 w 0 d LB B REP EGPN EG 0.110Japanesequail4 FD 12 w NR NR LB F REP PROG WO 0.194Chicken 5 FD 10 w NR NR LB F REP PROG WO 3.26Ringedturtle dove2 DR 11 w NR NR AD M REP TEWT TE 11.8Japanesequail2 FD 1 w 14 w JV F REP TPRD WO 93.1Japanesequail2 FD 27 d NR NR AD F REP PROG WO 377AX7-67Table AX7-1.3.3 (cont’d). Avian Toxicity Data Used to Develop the Eco-SSLAvianSpeciesNo. ofDosesRoute ofExposureExposureDurationDurationUnitsAgeAgeUnitsLifestage SexEffectTypeEffectMeasureResponseSiteNOAEL(mg/kgbw/day)LOAEL(mg/kgbw/day)GrowthJapanesequail3 FD 5 w 1 d JV F GRO BDWT WO 1.56 15.6Japanesequail3 FD 2 w 1 d JV B GRO BDWT WO 2.77Japanesequail2 FD 2 w 1 d JV NR GRO BDWT WO 4.64Japanesequail3 FD 4 w 0 d JV F GRO BDWT WO 5.93 59.3Chicken 4 FD 4 w 4 w JV NR GRO BDWT WO 6.14 61.4Chicken 4 FD 4 w 4 w JV NR GRO BDWT WO 7.10 71.0Japanesequail5 FD 12 w 0 d JV F GRO BDWT WO 11.1 111Japanesequail5 FD 12 w 1 w JV F GRO BDWT WO 11.2 112Japanesequail5 FD 2 w 6 d JV NR GRO BDWT WO 12.6 126Japanesequail5 FD 1 w 1 d JV NR GRO BDWT WO 13.5 67.4Chicken 2 FD 21 d 1 d JV B GRO BDWT WO 14.2Duck 3 GV 3 mo 24 w MA F GRO BDWT WO 20.0Americankestrel4 GV 10 d 1 d JV NR GRO BDWT WO 25.0 125Chicken 2 FD 20 d 1 d JV B GRO BDWT WO 28.4Japanesequail5 FD 14 d 1 d JV B GRO BDWT WO 34.5AX7-68Table AX7-1.3.3 (cont’d). Avian Toxicity Data Used to Develop the Eco-SSLAvianSpeciesNo. ofDosesRoute ofExposureExposureDurationDurationUnitsAgeAgeUnitsLifestage SexEffectTypeEffectMeasureResponseSiteNOAEL(mg/kgbw/day)LOAEL(mg/kgbw/day)Americankestrel4 FD 60 d 1-2 yr AD B GRO BDWT WO 54.3Chicken 5 FD 2 w 1 d JV M GRO BDWT WO 61.3 123Mallard 4 FD 8 d 9 d JV NR GRO BDWT WO 66.9Chicken 5 FD 20 d 1 d JV M GRO BDWT WO 38.2Chicken 2 FD 3 w 1 d JV M GRO BDWT WO 53.1Japanesequail3 FD 32 d NR NR AD F GRO BDWT WO 64.3Chicken 2 FD 19 d 1 d JV M GRO BDWT WO 76.3Chicken 3 FD 2 w 1 d JV M GRO BDWT WO 124Chicken 4 FD 14 d 8 d JV M GRO BDWT WO 152Chicken 2 FD 20 d 1 d JV M GRO BDWT WO 163Chicken 2 OR 4 w NR NR JV B GRO BDWT WO 200Chicken 2 FD 7 d 1 d JV M GRO BDWT WO 262Chicken 2 FD 2 w 1 d JV M GRO BDWT WO 270Chicken 2 FD 7 d 1 d IM NR GRO BDWT WO 273Chicken 2 FD 14 d 8 d JV M GRO BDWT WO 282AD = adult; ALWT = albumin weight; B = both; BDWT = body weight changes; d = days; DR = drinking water; EG = egg; EGG = effectson eggs; EGPN = egg production; ESTH = eggshellthinning; F = female; FD = food; GRO = growth; GV = gavage; JV =juvenile; LB = laying bird; MA = mature; M = male; mo = months; NR = not reported; OR = other oral; PROG = progeny countsor numbers; REP = reproduction; RSUC = reproductive success; SM = sexually mature; TE= testes; TEWT = testes weight; TPRD = total production; w = weeks; WO =whole organism; yr = years.Source: U.S. Environmental Protection Agency (2005b)AX7-69 Figure AX7-1.3.1. Avian reproduction and gr owth toxicity data considered in development of the Eco-SSL.

Source: U.S. Environmental Protection Agency (2005b).

Environmental Protection Agency, 2005b). For more information on the rationale for selecting TRVs, please refer to U.S. Environmental Protection Agency (2003). Many of the toxicity data presented in the Eco-SSL document (U.S. Environmental Protection Agency, 2005b) are lower than those discussed in the 1986 Lead AQCD. The TRV and resulting Eco-SSL were derived using many conservative assumptions. For example, the EPA (U.S. Environmental Protecti on Agency, 2005b) recognizes that toxicity is observed over a wide range of doses (<1 to >100 mg Pb/kg bw/d ay), even when considering only reproductive effects in the same species. In addition, the TR V of 1.63 mg/kg-day is lower than most of the reported doses that have been associated with measured effects. This is true for not only survival, growth, and reproductive effects but also biochemical, behavi oral, physiological, and pathological effects, which gene rally are observed at lo wer concentrations than effects on growth or reproduction. In addition, the Eco-SSL was back-calculated using conservative modeling assumptions. Therefore, the Eco-SSL of 11 mg/ kg may be considered a conservative value.AX7-70 Very little research has been done to expand the knowledge of the toxicity of Pb to birds since the Eco-SSL work was done. However, severa l studies have been c onducted on waterfowl. Toxicity data for waterfowl (in particular, mallards) were included in the soil Eco-SSL development process (Table AX7-1.3.3), alt hough mallards may be more exposed to

contaminants in sediment than soil. Effects on waterfowl may vary depending on the form of Pb, characteristics of the sediment, the foraging st rategy of the species (w hich may vary during reproduction), and the nutritional st atus of the animal. Sediment is recognized as an important route of exposure for waterfowl, particularly those species that da bble (i.e., forage on invertebrates in the sediment) (Beyer et al ., 2000; Douglas-Stroebel etal., 2005). DouglasStroebel et al. (2005) found that mallard ducklin gs exposed to Pb-contaminated sediment and a low nutrition diet exhibited more changes in be havior (as measured by time bathing, resting, and feeding) than Pb exposure or low nutrition exposu re alone. These effects may be due to the low nutrition diet being deficient in le vels of protein, amino acids, calci um, zinc, and other nutrients. Beyer et al. (2000) related blood Pb to subl ethal effects in wate rfowl along the Coeur

d = Alene River near a mining site in Idaho. The authors suggeste d that 0.20 mg/kg ww blood Pb represents the no-effect level. This no-effect blood concentra tion corresponds to a sediment Pb concentration of 24 mg/kg. A sediment concentr ation of 530 mg/kg, associated with a blood Pb concentration of 0.68 mg/kg ww, is suggested to be the lowest-effect concentration. These

results are consistent with those of Scheuhamm er (1989) who found blood Pb concentrations of 0.18 μg/mL to 0.65 μg/mL in mallards corresponded to conditions associated with greater than normal exposure to Pb, but that that should not be considered Pb poisoning. The study by Beyer et al. (2000) related blood Pb to waterfowl mortality and concl uded that some swan mortality may occur at blood Pb levels of 1.9 mg/kg ww, co rresponding to a sediment Pb concentration of 1800 mg/kg. Using the mean blood level of 3.6 mg/kg ww from all moribund swans in the study, it was predicted that half of the swans c onsuming sediment at th e 90th percentile rate would die with chronic exposure to sedi ment concentrations of 3600 mg/kg.

Mammalian Consumers

Effects on mammals observed in studies conducted since the 1986 Lead AQCD (U.S. Environmental Protection Agency, 1986a) are simila r to those reported previously: mortality, effects on reproduction, developmental effects, and changes in growth (U.S. EnvironmentalAX7-71 Protection Agency, 2005b). Very little research has been done to expand the knowledge of the toxicity of Pb to mammalian wildlife, sin ce the Eco-SSL work was done. Most studies conducted on mammals use laboratory animals to study potential adverse effects of concern for humans, and such studies are summarized in other sections of this document. Of the 2,429 papers identified in the literature search fo r Eco-SSL development, 219 met

the minimum criteria for further considera tion for mammalian Eco-SSL development (U.S. Environmental Protection Agency, 2005b). Th e 343 ecotoxicological endpoints for mammals

that were further evaluated included bioche mical, behavioral, physiological, pathological, reproductive, growth, and survival effects. Grow th and reproduction data were used to derive the Eco-SSL (Table AX7-1.3.4, Figure AX7-1.3.2). The geometric mean of the NOAELs was calculated as 40.7 mg/kg-day, wh ich was higher than the lowe st bounded LOAEL for survival, growth, or reproduction. Theref ore, the highest bounded NOAEL th at was lower than the lowest bounded LOAEL for survival, growth, or reproduc tion (4.7 mg Pb/kg bw-day) was used as the TRV (U.S. Environmental Protec tion Agency, 2005b). The TRV was used to back-calculate the Eco-SSL of 56 mg/kg soil (U.S. Environmen tal Protection Agency, 2005b). For more information on the rationale for selecting TRVs, please refer to U.S. Environmental Protection Agency (2003). A review of the data presented in the Ec o-SSL document (U.S. Environmental Protection Agency, 2005b) reveals that effects on survival generally are obser ved at Pb doses much greater than those reported in the 1986 Lead AQCD, wher e it was concluded that most animals would die when consuming a regular dose of 2 to 8 mg Pb/kg bw-day (U.S. Environmental Protection Agency, 1986a). However, the data presented in the Eco-SSL document (U.S. Environmental Protection Agency, 2005b) generally do not support this. While five studies reported decreased survival at these levels, 34 ot her studies reported no mortalit y or a LOAEL for mortality at significantly higher doses (U.S. Environmental Pr otection Agency, 2005b). Th e five studies that supported this low toxic level we re conducted on three species (m ouse, rat, and cow) and used either gavage or drinking wate r as the exposure method. The 34 other studies included data on these three species as well as five other spec ies (rabbit, dog, pig, hamster, and shrew) and included gavage and drinking water as well as food ingestion exposure methods. The NOAELs for survival ranged from 3.5 to 3200 mg/kg-da y (U.S. Environmental Protection Agency, 2005b). Therefore, the review of data in the Eco-SSL document s uggests effects on survival ofAX7-72Table AX7-1.3.4. Mammalian Toxicity Data Used to Develop the Eco-SSLMammalianSpeciesNo. ofDosesRoute ofExposureExposureDurationDurationUnits AgeAgeUnits Lifestage SexEffectTypeEffectMeasureResponseSiteNOAEL(mg/kgbw/day)LOAEL(mg/kgbw/day)ReproductionRat 5 DR 62 d 21 d GE F REP PRWT WO 0.71 7.00Rat 6 DR 21 d NR NR GE F REP PRWT WO 1.00 5.00Rat 3 DR 35 d NR NR AD M REP RSUC WO 2.60 26.0Rat 4 DR 62 d 21 d GE B REP PRWT WO 3.00 6.0Sheep 3 FD 27 w NR NR GE F REP RSUC WO 4.50 —Rat 6 DR 21 d NR NR GE F REP PRWT WO 5.00 10.0Guinea pig 3 DR 40 d NR NR GE F REP PRWT WO 5.50 —Rat 5 FD 92 w 21 d JV M REP TEWT TE 7.50 74.9Rat 4 DR 23.8 d 21 d LC F REP Other WO 8.90Rat 5 DR 23.8 d 21 d GE F REP Other WO 9.10 45.0Cotton rat 3 DR 7 w NR NR AD M REP RHIS RT 12.4 170Rat 4 GV 9 w 10 w JV M REP SPCV TE 18.0 180Rat 3 DR 100 d 21 d GE F REP PRWT WO 25.4 —Rat 2 FD 35 d 70 d LC F REP PRWT WO 27.5 —Rat 4 DR 60 d NR NR SM M REP TEWT TE 31.6 63.2Rat 4 DR 56 d 70 d LC F REP PROG WO 32.5 —Rat 3 DR 31 d NR NR LC F REP PRWT WO 33.3 111Rat 4 GV 41 d NR NR GE F REP PRWT WO 41.0 54.6Rat 5 DR 1 w 94 d JV M REP SPCL SM 47.3 82.0Rat 4 DR 30 d NR NR SM M REP Other SV 56.0 285AX7-73Table AX7-1.3.4 (cont’d). Mammalian Toxicity Data Used to Develop the Eco-SSLMammalianSpeciesNo. ofDosesRoute ofExposureExposureDurationDurationUnitsAgeAgeUnitsLifestage SexEffectTypeEffectMeasureResponseSiteNOAEL(mg/kgbw/day)LOAEL(mg/kgbw/day)Hamster 2 DR 51 d 15 w GE F REP PROG WO 64.8 —Hamster 2 DR 14 d 11 w GE F REP PROG WO 64.9 —Rat 4 DR 37 d NR NR GE F REP PRWT WO 90.1 270Rat 5 GV 12 d NR NR GE F REP RSEM EM 100 150Rat 3 DR 68 d 25 d GE F REP PRWT WO 115 —Rat 4 DR 77 d 25 d GE F REP PRWT WO 116 —Rat 2 DR 21 d NR NR LC F REP PRWT WO 120 —Mouse 3 FD 8 w 2 mo GE M REP SPCV TE 144 1,440Mouse 7 FD 30 d NR NR LC F REP PRWT WO 202 506Mouse 7 FD 30 d NR NR LC F REP PRWT WO 202 506Rat 4 DR 21 d NR NR GE F REP DEYO WO 276 552Rat 5 DR 10 w NR NR AD M REP TEWT MT 294 587Rat 2 GV 102 d 30 d GE F REP PRWT WO 441 —Rat 2 DR 9 mo NR NR SM M REP RHIS TE 600 —Rat 4 FD 4 d NR NR LC F REP PRWT WO 601 1,500Rat 4 DR 13 w NR NR JV M REP FERT WO 639 —Mouse 4 GV 60 d NR NR AD F REP RPRD OV — 2.00Rat 3 FD 339 d 26-27 d JV B REP PRWT WO — 2.49Rat 2 DR 9 mo 21 d JV F REP DEYO WO — 2.94Mouse 2 DR 6 mo 21 d JV F REP DEYO WO — 3.62Mouse 4 GV 52 d 2 mo GE F REP PROG EM — 5.50Rat 2 DR 120 d 1 d GE M REP SPCL TE — 6.76AX7-74Table AX7-1.3.4 (cont’d). Mammalian Toxicity Data Used to Develop the Eco-SSLMammalianSpeciesNo. ofDosesRoute ofExposureExposureDurationDurationUnitsAgeAgeUnitsLifestage SexEffectTypeEffectMeasureResponseSiteNOAEL(mg/kgbw/day)LOAEL(mg/kgbw/day)Mouse 2 DR 5 w NR NR AD M REP TEDG TE — 16.6Mouse 2 GV 2 w NR NR JV M REP SPCL SM — 46.4Rat 2 FD 102 d NR NR GE F REP PROG WO — 49.6Rat 2 GV 3 mo 8 w SM M REP TEDG TE — 50.0Rat 2 DR 18 d NR NR GE F REP PRWT WO — 55.5Rat 3 DR 90 d NR NR AD M REP SPCL SM — 61.2Mouse 2 DR 23 d NR NR GE F REP PRWT WO — 78.6Mouse 2 DR 62 d NR NR GE F REP PRWT WO — 99.8Mouse 2 DR 18 w 6-8 w LC F REP PRWT WO — 137Mouse 2 DR 12 w 9 w SM M REP PRFM WO — 139Mouse 4 FD 18 d NR NR GE F REP PRWT WO — 154Rat 2 DR 4 w 99 d JV M REP SPCL SM — 171Rat 5 DR 6 w 4 mo GE F REP RHIS WO — 175Rat 2 DR 22 d NR NR GE F REP PRWT WO — 178Rat 3 DR 30 d 52 d JV M REP GREP PG — 198Rat 2 DR 13 w NR NR GE F REP PRWT WO — 200Rat 2 DR 21 d 80 d JV F REP PRWT WO — 218Rat 4 FD 3 w NR NR LC F REP PRWT WO — 221Rat 2 FD 1 w 19 w LC F REP PRWT WO — 222Rat 4 FD 3 w NR NR LC F REP PRWT WO — 230Rat 3 FD 25 d NR NR LC F REP PRWT WO — 258Rat 2 DR 21 d NR NR LC F REP PRWT WO — 330AX7-75Table AX7-1.3.4 (cont’d). Mammalian Toxicity Data Used to Develop the Eco-SSLMammalianSpeciesNo. ofDosesRoute ofExposureExposureDurationDurationUnitsAgeAgeUnitsLifestage SexEffectTypeEffectMeasureResponseSiteNOAEL(mg/kgbw/day)LOAEL(mg/kgbw/day)Rat 2 DR 30 d 52 d JV M REP SPCL SM — 354Rat 2 DR 17 d NR NR GE F REP PRWT WO — 360Rat 2 DR 24 d NR NR LC F REP PRWT WO — 360Rat 2 DR 12 d NR NR GE F REP PRWT WO — 362Rat 2 DR 30 d 27 d JV M REP SPCL SM — 364Mouse 2 DR 44 d NR NR GE F REP PRWT WO — 381Mouse 2 DR 14 d NR NR LC F REP PRWT WO — 381Rat 2 DR 50 d 24 d JV F REP RBEH WO — 381Mouse 2 DR 45 d 50-100 d GE F REP ODVP WO — 404Rat 2 DR 22 d NR NR GE F REP PRWT WO — 420Mouse 2 DR 48 d NR NR GE F REP PRWT WO — 437Rat 2 DR 9 mo 3 mo SM M REP SPCL TE — 579Rat 2 DR 9 mo NR NR SM M REP TEDG TE — 600Rat 2 DR 3 w 14 w LC F REP PRWT WO — 635Mouse 2 FD 7 d NR NR GE F REP RSUC EM — 646Rat 2 DR 126 d 1 d GE F REP PROG WO — 651Rat 2 DR 20 w 10 w GE F REP PRWT WO — 750Mouse 2 DR 4 d NR NR LC F REP PRWT WO — 762Rat 2 FD 2 w NR NR LC F REP PRWT WO — 828Rat 2 FD 7 d NR NR LC F REP PRWT WO — 833Rat 2 FD 21 d NR NR LC F REP PRWT WO — 991Mouse 4 DR 18 w 11 w JV F REP TEWT WO — 1,370AX7-76Table AX7-1.3.4 (cont’d). Mammalian Toxicity Data Used to Develop the Eco-SSLMammalianSpeciesNo. ofDosesRoute ofExposureExposureDurationDurationUnitsAgeAgeUnitsLifestage SexEffectTypeEffectMeasureResponseSiteNOAEL(mg/kgbw/day)LOAEL(mg/kgbw/day)Rat 2 FD 30 d NR NR LC F REP PRWT WO — 1,770Mouse 2 DR 14 w NR NR GE B REP PROG WO — 1,990Rat 2 FD 16 d NR NR LC F REP PROG WO — 2,570Rat 2 FD 7 d NR NR LC F REP PRWT WO — 2,570Rat 2 FD 25 d NR NR LC F REP PRWT WO — 2,570Rat M FD 27 d NR NR LC C REP PROG WO — 2,840Mouse 2 DR 14 w 21 d JV B REP PROG WO — 3,630Rat 2 FD 17 d NR NR LC F REP PRWT WO — 6,170GrowthHorse 2 FD 15 w 20-21 w JV M GRO BDWT WO 0.15 —Rat 2 FD 21 d 0 d JV F GRO BDWT WO 0.5 —Rat 6 DR 21 d NR NR GE F GRO BDWT WO 1.00 5.00Rat 5 DR 7 d 50 d AD F GRO BDWT WO 1.27 13.0Cattle 4 OR 7 w 1 w JV M GRO BDWT WO 1.99 —Rat 3 DR 14 d 21 d JV F GRO BDWT WO 2.40 —Rat 2 DR 332 d 28 d JV B GRO BDWT WO 2.98 —Rat 4 DR 7 w 21 d GE F GRO BDWT WO 4.70 8.90Dog 3 FD 7 mo NR NR JV NR GRO BDWT WO 4.71 —Rat 3 DR 30 d 22-24 d JV M GRO BDWT WO 5.64 28.2Rat 4 DR 23 d 22 d JV F GRO BDWT WO 5.80 29.0Cattle 3 OR 84 d NR NR JV M GRO BDWT WO 7.79 —Rat 2 OR 6 w NR NR AD M GRO BDWT WO 9.10 —AX7-77Table AX7-1.3.4 (cont’d). Mammalian Toxicity Data Used to Develop the Eco-SSLMammalianSpeciesNo. ofDosesRoute ofExposureExposureDurationDurationUnitsAgeAgeUnitsLifestage SexEffectTypeEffectMeasureResponseSiteNOAEL(mg/kgbw/day)LOAEL(mg/kgbw/day)Rat 2 GV 8 w NR NR JV F GRO BDWT WO 10.0 —Rat 3 DR 6 mo NR NR AD M GRO BDWT WO 10.6 532Rabbit 3 GV 10 d 1 d JV F GRO BDWT WO 10.7 50.4Rat 2 DR 140 d 21 d JV M GRO BDWT WO 10.7 —Rat 2 DR 6 w NR NR JV M GRO BDWT WO 15.1 —Rat 2 FD 10 w NR NR JV M GRO BDWT WO 15.4 —Rat 2 OR 6 w NR NR AD M GRO BDWT WO 15.5 —Rat 2 DR 7 w NR NR JV M GRO BDWT WO 16.1 —Mouse 3 DR 14 d 0 d JV NR GRO BDWT WO 16.3 163Rat 4 GV 9 w 10 w JV M GRO BDWT WO 18.0 180Rat 3 FD 339 d 26-27 d JV B GRO BDWT WO 18.3 —Rat 4 GV 29 d NR NR SM F GRO BDWT WO 18.9 —Rat 7 DR 10 w NR NR JV M GRO BDWT WO 24.3 —Rat 4 DR 56 d 70 d LC F GRO BDWT WO 32.5 —Sheep 5 FD 84 d NR NR JV M GRO BDWT WO 32.7 —Rat 2 DR 10 w NR NR JV M GRO BDWT WO 38.5 —Cattle 4 FD 7 w 16 w JV M GRO BDWT WO 43.0 —Rat 2 GV 28 d 2 d JV B GRO BDWT WO 50.0 —Rat 5 DR 4 w 94 d JV M GRO BDWT WO 71.5 178Rat 4 GV 12 d 2 d JV B GRO BDWT WO 75.0 225Rat 2 FD 4 w NR NR JV M GRO BDWT WO 100 —Rat 6 DR 10 w NR NR JV M GRO BDWT WO 120 383AX7-78Table AX7-1.3.4 (cont’d). Mammalian Toxicity Data Used to Develop the Eco-SSLMammalianSpeciesNo. ofDosesRoute ofExposureExposureDurationDurationUnitsAgeAgeUnitsLifestage SexEffectTypeEffectMeasureResponseSiteNOAEL(mg/kgbw/day)LOAEL(mg/kgbw/day)Mouse 3 FD 4 w 3 mo JV B GRO BDWT WO 136 1360Mouse 2 DR 18 w 6-8 w LC F GRO BDWT WO 137 —Mouse 2 DR 12 w NR NR GE M GRO BDWT WO 139 —Rat 3 DR 30 d 52 d JV M GRO BDWT WO 169 508Rat 2 DR 4 w 99 d JV B GRO BDWT WO 171 —Rat 4 GV 18 d 3 d JV M GRO BDWT WO 180 —Mouse 3 DR 6 w 7 w SM M GRO BDWT WO 187 373Rat 4 GV 18 d 2 d JV B GRO BDWT WO 200 —Rat 2 GV 91 d NR NR JV M GRO BDWT WO 200 —Rat 2 DR 21 d 80 d JV F GRO BDWT WO 218 —Rat 4 FD 1 w NR NR LC F GRO BDWT WO 230 460Rat 4 DR 30 d NR NR JV M GRO BDWT WO 285 —Mouse 5 DR 10 w NR NR JV M GRO BDWT WO 362 —Rat 2 DR 30 d 52 d JV M GRO BDWT WO 364 —Rat 4 GV 14 d 14 d JV NR GRO BDWT WO 400 800Rat 5 GV 14 d 20 d JV NR GRO BDWT WO 400 800Rat 2 FD 14 mo 0 d JV NR GRO BDWT WO 431 —Rat 2 GV 102 d 30 d LC F GRO BDWT WO 441 —Mouse 4 GV 12 d 6 d JV M GRO BDWT WO 534 —Mouse 7 FD 30 d NR NR LC F GRO BDWT WO 632 1264Rat 2 DR 126 d 1 d GE F GRO BDWT WO 651 —Rat 2 DR 20 w 10 w GE F GRO BDWT WO 750 —AX7-79Table AX7-1.3.4 (cont’d). Mammalian Toxicity Data Used to Develop the Eco-SSLMammalianSpeciesNo. ofDosesRoute ofExposureExposureDurationDurationUnitsAgeAgeUnitsLifestage SexEffectTypeEffectMeasureResponseSiteNOAEL(mg/kgbw/day)LOAEL(mg/kgbw/day)Mouse 7 FD 28 d NR NR LC F GRO BDWT WO 1260 2530Rat 4 FD 18 d NR NR LC F GRO BDWT WO 1500Rat 2 DR 9 d 21 d JV M GRO BDWT WO — 3.30Cattle 2 FD 283 d 7 mo JV M GRO BDWT WO — 15.0Rat 3 DR 92 d 25 d GE F MPH GMPH TB — 28.7Rat 4 DR 7 d 25 d GE F GRO BDWT WO — 29.0Rat 2 DR 5 d 26 d JV F GRO BDWT WO — 29.0Rat 2 DR 26 d 22 d JV F GRO BDWT WO — 29.5Rat 2 DR 14 d 26 d JV F MPH Other TA — 29.9Rat 2 DR 10 d 26 d JV F GRO BDWT WO — 30.4Mouse 2 GV 3 w NR NR JV M GRO BDWT WO — 46.4Dog 2 OR 5 w <1 yr JV NR GRO BDWT WO — 50.0Shrew 4 FD 31 d NR NR JV B GRO BDWT WO — 61.5Rat 3 GV 58 d 2 d JV B GRO BDWT WO — 100Pig 2 FD 13 w 4 w JV NR GRO BDWT WO — 173Rat 2 GV 29 d 2 d JV F GRO BDWT WO — 200Rat 2 FD 5 w NR NR MA NR GRO BDWT WO — 272Rat 2 GV 6 d 1 d JV B GRO BDWT WO — 328Rat 2 DR 30 d 27 d JV M GRO BDWT WO — 354Rat 2 DR 50 d 24 d JV M GRO BDWT WO — 371Rat 2 GV 28 d 2 d JV M GRO BDWT WO — 400Rat 4 GV 14 d 18 d JV NR GRO BDWT WO — 400AX7-80Table AX7-1.3.4 (cont’d). Mammalian Toxicity Data Used to Develop the Eco-SSLMammalian SpeciesNo. ofDosesRoute ofExposureExposureDurationDurationUnitsAgeAgeUnitsLifestage SexEffectTypeEffectMeasureResponseSiteNOAEL(mg/kgbw/day)LOAEL(mg/kgbw/day)Mouse 2 DR 45 d 50-100 d GE F GRO BDWT WO — 404Rat 4 FD 1 w NR NR LC F GRO BDWT WO — 442Rat 2 DR 6 w 14 w LC F GRO BDWT WO — 638Mouse 4 DR 10 w 11 w JV F GRO BDWT WO — 748Rat 2 FD 21 d NR NR LC F GRO BDWT WO — 991Rat 2 GV 18 d 2 d JV B GRO BDWT WO — 1000Rat 2 FD 2 w 0 d JV NR GRO BDWT WO — 1430Rat 4 GV 14 d 24 d JV NR GRO BDWT WO — 1600Rat 2 FD 2 w 60-80 d JV M GRO BDWT WO — 2390Rat 3 GV 14 d 16 d JV NR GRO BDWT WO — 2400Rat 2 FD 14 d 60 d JV M GRO BDWT WO — 2650AD = adult; B = both; BDWT = body weight changes; d = days; DEYO= death of young; DR = drinking water; F = female; FD = food;FERT = fertility; GMPH = general morphology;GRO = growth; GV = gavage; JV = juvenile;LC = lactation; M = male; MA = mature; mo = months; MPH = morphology; NR = not reported; ODVP = offspring development; OR = other oral;PG = prostate gland; PROG = progeny counts or numbers; PRWT = progeny weight; RBPH = reproductive behavior; REP = reproduction;RHIS = reproductive organ histology; RSEM = resorbedembryos; RSUC = reproductive success (general); RT = reproductive tissue; SM = sexually mature; SPCL = sperm cell counts; SPCV= sperm cell viability; TA = tail; TB = tibia; TE = testes;TEDG = testes degeneration; TEWT= testes weight; w = weeks; WO = whole organism; yr = years.Source: U.S. Environmental Protection Agency (2005b).AX7-81 Figure AX7-1.3.2. Mammalian reproduction and growth toxicity data considered in development of the Eco-SSL.

Source: U.S. Environmental Protection Agency (2005b).

wildlife generally would occur at doses greater than the 2 to 8 mg /kg-day reported to be toxic to most animals in the 1986 Lead AQCD (U.S. Environmental Protection Agency, 1986a).

AX7.1.3.4 Recent Studies on the Eff ects of Lead on Decomposers Recent studies on effects of Pb to two groups of decomposers are summarized in this

subsection. Effects on terrestrial invertebrates, such as earthwor ms and springtails, are described first, followed by effects on microorganisms.

Effects on Invertebrates Since the 1986 Lead AQCD, there have been several studies in which terrestrial invertebrates were exposed to Pb in soil. Ma ny of these were reviewed during the development of the Eco-SSLs (U.S. Environmental Protecti on Agency, 2005b). The relevant information

from the Eco-SSL document is described below.AX7-82 A literature search and review was conducted to identify critical papers published since

2002. Effects on earthworms and other invertebra tes observed in stud ies conducted since the 1986 Lead AQCD are similar to those reported prev iously: mortality and decreased growth and reproduction (Lock and Janssen, 2002; Davies et al., 2002; Rao et al., 2003; Bongers et al., 2004; Nursita et al., 2005; U.S. Environmental Protection Agency, 2005b). The literature search completed for terres trial invertebrate Eco-SSL development identified 179 papers for detailed review, of which 13 met the minimum criteria for further consideration (U.S. Environmen tal Protection Agency, 2005b). Most of the 18 ecotoxicological endpoints that were further eval uated measured reproduction or survival as the ecologically relevant endpoint. Four of these, represen ting one species under th ree different pH test

conditions were used to de velop the Eco-SSL of 1700 mg/ kg soil (Table AX7-1.3.5).

Table AX7-1.3.5. Invertebrate Toxicity Data Used to Develop the Eco-SSL

Invertebrate Species Soil pH % Organic Matter Toxicity Parameter Pb in Soil (mg/kg dw) Collembola

(Folsomia candida ) 6.0 10 MATC 1 (reproduction) 3162

Collembola 4.5 10 MATC (reproduction) 3162

Collembola 5.0 10 MATC (reproduction) 894

Collembola 6.0 10 MATC (reproduction) 894

Geometric Mean 1682

* MATC = Maximum Acceptable Threshold Concentration, or the geometric mean of the NOEC (no-observedeffect concentration) an d LOEC (lowest-observed-effect concentration).

Source: U.S. Environmental Protection Agency (2005b).

In a study designed to test the t oxicity of Pb to the earthworm Eisenia fetida , Davies et al. (2002) found that the 28-day LC 50 (± 95% confidence intervals) for Pb in soils contaminated with Pb(NO 3 ) 2 was 4379 ∀ 356 mg/kg. Twenty-eight day EC 50 values (±95% confidence intervals) for weight change and cocoon production were 1408 ∀ 198 andAX7-83 971 ∀ 633 mg/kg, respectively. Significant mortalit ies were noted at concentrations of 2000 mg/kg. These data are consistent with t hose reported in the Eco-SSL document (U.S. Environmental Protection Agency, 2005b) fo r the same species of earthworm. Nursita et al. (2005) found no mortality a nd no adverse effects on reproduction (i.e., number of juveniles) of the collembolan Proisotoma minuta exposed for 42 days to 300, 750, 1500, or 3000 mg Pb/kg as Pb-nitrate in an acidic (pH = 4.88) sandy loam soil. It was noted that the soils were allowed to equilibrate for 4 weeks after adding the Pb-nitrate before the organisms were added. The observation of no effect at 3000 mg /kg is consistent with that of Sandifer and Hopkin (1996). Sandifer and Hopkin (1996) de termined a NOEC (no-observed-effect concentration) and LOEC (lowest-observed-eff ect concentration) for collembolan reproduction of 2000 and 5000 mg/kg, respectively. (A MATC [maximum-acceptable-threshold concentration] of 3162 mg/kg was us ed to develop the Eco-SSL). The remaining 14 toxicity endpoints that we re not used to develop the Eco-SSL for invertebrates are presented in Table AX7-1.3.6. None of these e ndpoints was considered eligible

for Eco-SSL derivation. Lock and Janssen (2002) exposed the potworm Enchytraeus albidus to Pb, as Pb-nitrate. The 21-day LC 50 was 4530 mg/kg, and the 42-day EC 50 for juvenile reproduction was 320 mg/kg. The F1 generation was then grown to maturity in the same concentration soil and subsequently used in a reproduction test. The EC 50 for the F1 generation (394 mg/kg) was similar to that of the P generation. The author s concluded that the two- generation assay did not increase the sensitivity of the test (Lock and Janssen, 2002). None of the 18 toxicity endpoints evaluated in detail during development of the Eco-SSLs used this species. The LC 50 reported for the potworm was higher than reported for nema todes and similar to that reported for the

earthworm. The EC 50 for reproduction was lower than repor ted for the earthworm or collembola. Recent work by Bongers et al. (2004) cautioned ag ainst attributing all toxicity observed in a spiked-soil toxicity test to Pb. They found that the counterion may also contribute to the toxicity of Pb in the springtail Folsomia candida . This may have implications on the interpretation of the Eco-SSL data, because the t oxicity of the counterion (nitrate) was not taken into account during Eco-SSL development. Pe rcolation (removal of the counterion) had no statistically significant effect on Pb-chloride toxicity (LC 50 = 2900 mg/kg for both nonpercolated and percolated soil; EC 50 for reproduction = 1900 mg /kg or 2400 mg/kg forAX7-84 Table AX7-1.3.6. Invertebrate Toxicity Da ta Not Used to Develop the Eco-SSL

Invertebrate Species Soil pH % Organic Matter Toxicity Parameter Pb in Soil (mg/kg dw)
Nematode 4 1.14 LC 50 285

Nematode 4 1.14 LC 50 297

Nematode 4 4.2 LC 50 847

Nematode 4 4.2 LC 50 1341

Nematode 6.2 1.7 LC 50 1554

Nematode 5.1 3.0 LC 50 891

Earthworm 6.3 10.0 EC 50 1940

Earthworm 6.1 10.0 EC 50 1629

Earthworm 6.0 10.0 LC 50 3716

Earthworm 6.5 10.0 ILL 1.16

Nematode 4 10 LC 50 1434

Nematode 4 10 NOAEC 2235

Nematode 6.1 3.4 LC 50 13.9

Nematode 6.2 2.2 LC 50 11.6

*NOAEC (no-observed-adverse-effect concentration); LC 50 (concentration lethal to 50% of test population); EC 50 (effect concentration for 50% of test po pulation); ILL (incipient lethal level).

Source: U.S. Environmental Protection Agency (2005b).

non-percolated or percolated soil, respectively). However, percolation did have a significant effect on Pb-nitrate toxicity (LC 50 = 980 mg/kg or 2200 mg/kg for non- percolated or percolated soil, respectively; EC 50 for reproduction = 580 mg/kg or 1700 mg/kg for non-percolated or percolated soil, respectively). Lead nitrate was more toxic th an Pb-chloride for survival and reproduction. However, the toxicity of Pb, from chloride or nitrate, was not significantly different after the counterion was perc olated out of the test soil. It is noted that the soil was left for 3 weeks to equilibrate before testing. Lock and Janssen (2002) also found that Pb-nitrate was more toxic than Pb-chloride, and they used Pb -nitrate in their experiments because 1000 mg/kg Pb-chloride did not produce any mortality in their range-finding tests. This difference inAX7-85 chloride and nitrate toxicity has not been found for earthworms (Neuhauser et al., 1985; Bongers et al., 2004). Rao et al. (2003) exposed the earthworm Eisenia fetida to Pb-oxide in an artificial soil with a pH of 6 at the LC 50 concentration of 11 mg/kg. Exposure for 14 days resulted in a number of effects including body fragmentat ion, protrusions, rupture of the cuticle, etc. Many of these effects may trigger defensive mechanisms. For ex ample, fragmentation of the affected posterior region was followed by regeneration and a new ect oderm layer was formed to cover affected areas, both of which processes may serve to pr event soil bacteria from further affecting the earthworm (Rao et al., 2003).

Effects on Microorganisms and Microbial Processes Microorganisms and microbial processes were not included in the Eco-SSL development process (see Attachment 1-2 of OSWER Di rective 92857-55 dated November 2003 in U.S. Environmental Protection Agency [2005a]). Many reasons were given, including that it is unlikely that site conditions would only pose unaccep table risk to microbes and not be reflected as unacceptable risks to higher organisms; that th e significance of laborator y-derived effects data to the ecosystem is uncertain; and that the sp atial (across millimeter distances) and temporal (within minutes to hours) vari ation makes understanding ecologica l consequences challenging. Microbial endpoints often vary dramatically based on moisture, temperature, oxygen, and many

non-contaminant factors. Th erefore, the recommendation arising from the Eco-SSL development process was that risks to microbes or microbial processes not be addressed through the chemical screening process bu t that they should be addresse d within a site-specific risk assessment (U.S. Environmental Protection Agency, 2005b).

Few studies on the effects of Pb to microbial processes have been published since 1986. As the direct toxicity to fungi and bacterial popu lations are difficult to determine and interpret,

indicators for soil commun ities are often measured as proxies for toxicity (e.g., ur ease activity in soil). Recent studies of this nature (Doelman and Haanstra, 1986; Wilke, 1989; Haanstra and Doelman, 1991) are summarized in this subsecti on. The Pb concentrations in these recent studies (1000 to 5000 mg/kg) are consistent with thos e reported in the 1986 Lead AQCD (U.S. Environmental Protection Agency, 1986a) as asso ciated with effects on microbial processes (750 to 7500 mg/kg).AX7-86 The effects of Pb-chloride on the processes of nitrification and n itrogen mineralization were studied in a 28-day experiment by Wilke (1 989). The authors reported that nitrification was increased by 12 and 16% at levels of 1000 and 4000 mg/kg, respectively, and that nitrogen mineralization was reduced by 32 and 44% at concentrations of 1000 and 4000 mg/kg, respectively. The effects of Pb on arylsulfatase (Haanst ra and Doelman, 1991) and urease activity (Doelman and Haanstra, 1986) in soil were investigated. LC 50 s for decreases in arylsulfatase activity were reported at Pb concentrations of 3004 and 4538 mg/kg in a silty loam soil, at pH 6 and 8, respectively. The LC 50 for a decrease in urease activ ity was 5060 mg Pb/kg in a sandy

loam soil. In laboratory microcosm studies Cotrufo et al. (1995) found that decomposition of oak ( Quercus ilex ) leaf litter was reduced at elevated Pb (~20 mg Pb g ! 1 C) levels after 8 months compared to controls (~2 mg Pb g ! 1 C). The researchers found soil respiration and amount of soil mycelia correlated negatively with soil Pb, Zn and Cr concentration.

AX7.1.3.5 Summary The current document expands upon and updates knowledge related to the effects of Pb on terrestrial primary producers, consumers, and decomposers.

Primary Producers The effects of Pb on terrestrial plants incl ude decreased photosynthe tic and transpiration rates in addition to decreased gr owth and yield. The phytotoxicity of Pb is considered to be relatively low, and there are few reports of phytotoxicity from Pb exposure under field conditions. Recently, phytotoxicity data were reviewed for the development of the Eco-SSL (U.S. Environmental Protection Agency, 2005b). Many of the toxicity data presented in the EcoSSL document (U.S. Environmental Protection Agen cy, 2005b) are lower than those discussed in the 1986 Lead AQCD, although both documents acknowle dged that toxicity is observed over a wide range of concentrations of Pb in soil (tens to thousands of mg/kg soil). This may be due to many factors, such as soil conditions (e.g., pH, orga nic matter) and differences in bioavailability of the Pb in spiked soils perhaps due to lack of equilibration of the Pb solution with the soil after spiking. Most phytotoxicity data continue to be developed for agricultu ral plant species (i.e.,AX7-87 vegetable and grain crops). Few data are av ailable for trees or na tive herbaceous plants, although two of the five toxicity endpoints used to develop the Eco-SSL were for trees and two

were for clover.

Consumers

Effects of Pb on avian and mammalian consumers include decreased survival, reproduction, and growth as well as effects on development and behavior. There remain few field effects data for consumers, except from sites with multiple contaminants, for which it is difficult to attribute toxicity specifically to Pb. Avian and mammalian toxicity data recently were reviewed for the development of Eco- SSLs (U.S. Environmental Protection Agency, 2005b). Many of the toxicity data presented in the Eco-SSL document (U.S. Environmental Protection Agency, 2005b) are lower than those discussed in the 1986 Lead AQCD, although the EPA (U.S. Environmental Protecti on Agency, 2005b) recognizes that toxicity is observed over a wide range of doses (<1 to >1000 mg Pb/kg bw-day ). Most toxicity data for birds have been derived from chicken and quail studies, and most data for mammals have been derived from laboratory rat and mouse studies. Data derive d for other species woul d contribute to the understanding of Pb toxicity, particularly for wi ldlife species with different gut physiologies. In addition, data derived using environmentally realistic exposures, such as from Pbcontaminated soil and food, may be recommende d. Finally, data derived from inhalation exposures, which evaluate endpoints such as survival, growth, and reproduction, would contribute to understanding the implica tions of airborne releases of Pb.

Decomposers

Effects of Pb on soil invertebrates include decreased survival, growth, and reproduction. Effects on microorganisms include changes in nitrogen mineraliza tion and enzyme activities. Recent data on 1986 Lead toxicity to soil inverteb rates and microorganisms are consistent with those reported in the 1986 L ead AQCD (U.S. Environmental Protection Agency, 1986a), with toxicity generally observed at concentrations of hundreds to thousands of mg/kg soil. Studies on microbial processes may be influenced signifi cantly by soil parameters, and the significance of

the test results is not clear.AX7-88 Ecological Soil Screening Levels (Eco-SSLs)

Eco-SSLs are concentrations of contaminants in soils that would result in little or no measurable effect on ecological receptors (U.S. Environmental Protection Agency, 2005a). They were developed following rigorous scientific protocols and were s ubjected to two rounds of peer review. Due to conservative modeling assumptions (e.g., metal exists in most toxic form or highly bioavailable form, high food ingestion rate, high soil inges tion rate) which are common to screening processes, several Eco-SSLs are derived below the average background soil concentration for a particular contaminant. The Eco-SSLs for terrestrial plants, birds, mammals, and soil invertebrates are 120, 11, 56, and 1700 mg Pb/kg soil, respectively.



### AX7.1.4 Effects of Lead on Na tural Terrestrial Ecosystems The concept that organisms are part of larger systems that include both biotic and abiotic components of the environment date s back to the naturalis ts of the Victorian era. However, the breakthrough in what we now consider the ecosy stem approach to ecolo gy occurred in the 1950s and 1960s when E.P. and H.T. Odum pioneered the quantitative anal ysis of ecosystems (Odum, 1971). This approach encouraged th e calculation of energy flows into, out of, and within explicitly defined ecosystems. The ra pid development of com puter technology aided in the growth of ecosystem ecology by allowing the development and use of increasingly complex models for estimating fluxes that could not be directly measured. It was not long before the qua ntitative analysis of ecosystems was extended to examine the flows of nutrients and other chemical comp ounds. In temperate terrestrial systems, the

watershed was identified as a convenient and informative experimental unit (Bormann and Likens, 1967). A major conceptual breakthrough in the watershed approach was that drainage water chemistry could be used as an indicator of the “health” of the eco system. In a system limited by nitrogen, for example, elevated concentrations of NO 3 ! in drainage waters indicate that the ecosystem is no longer making optimal use of available nutrients. The ecosystem approach can also be used effectively in the st udy of trace element biogeochemistry. Input-output budg ets can be used to determine whether an ecosystem is a net source or sink of a trace element. Changes to th e input-output balance over time can be used to

assess the effects of natural or experimental changes in deposition, land use, climate, or otherAX7-89 factors. In addition, examinati on of fluxes within the ecosystem (in plant uptake, soil solutions, etc.) can be used to understand the processes that are most influen tial in determining the fate and transport of the trace element. Many published ecosystem studies include data for 1 to 3 years, the typical duration of research grant funding or doctoral dissertation research. While these studies enrich our understanding of terrestrial ecosystems, the most valuable studies are those that are maintained over many years. Natural variations in climate, pests, animal migrations, and other factors can

make inferences from short-term studies mi sleading (Likens, 1989). To nurture long-term research, the National Science Foundation s upports a network of Long-Term Ecological Research (LTER) sites that re present various biomes. This section describes terrestrial ecosystem research on Pb, focusing on work done since the 1986 Lead AQCD (U.S. Environmental Prot ection Agency, 1986a) and highlighting key long-term studies. Unfortunately, there are few studies that featur e long-term data on trace metal behavior at multiple levels of or ganization. Therefore, this exam ination of the effects of Pb on terrestrial ecosystems combines insights from long- and short-term investigations as well as studies at scales including whole ecosystems, communities, populations, and individual species.

AX7.1.4.1 Effects of Terrestrial Ecos ystem Stresses on Lead Cycling Terrestrial ecosystems may respond to stressors in a variety of ways, including reductions in the vigor and/or growth of vegetation, reduc tions in biodiversity, and effects on microbial processes. Each of these effects may lead to the “leakage” of nutrients, especially nitrogen, in drainage waters. The reduced vi gor or growth of vegetation resu lts in a lower uptake of nitrogen

and other nutrients from soils. Reduced biodi versity accompanied by lower total net primary productivity for the ecosystem would also result in a lower nutrient uptake. Effects of stress in microbial populations are less obv ious. If the stress reduces mi crobial activity rates, then nutrients bound in soil organic matter (e.g., organic nitrogen compounds) will likely be

mineralized at a lower rate and re tained in the system. On the other hand, disturbances such as clear-cutting, ice-storm damage, and soil freezing can result in substantial nutrient losses from soils (Bormann et al., 1968; Li kens et al., 1969; Mitchell et al., 1996; Groffman et al., 2001; Houlton et al., 2003).AX7-90 Since the movement and fate of Pb in terrest rial ecosystems is strongly related to the organic matter cycle (Section AX7.1.2), stressors that could lead to disrupti on or alteration of the soil organic matter pool are of part icular concern in assessing effects of ecosystem stress on Pb cycling. By binding soluble Pb, so il organic matter acts as a ba rrier to the release of Pb to drainage waters (Wang et al., 1995; Kaste et al., 2003; Watmough and Hutchinson, 2004). As a result, concentrations of Pb in soil solutions and drainage waters tend to be low (Driscoll et al., 1988; Wang et al., 1995; Bacon and Bain, 1995; Johnson et al., 1995a). Through decomposition and leaching, soluble organic matter is released to solution, and with it, some Pb is also mobilized. Wang and Benoit (1996) found that essentially all of the Pb in soil solutions in a hardwood forest in New Hampshire was bound to di ssolve organic matter (DOM). This release of soluble Pb does not typically result in elevated surface wate r Pb concentrations, because (1) organic matter has a relatively long residence time in most temperate soils (Gosz et al., 1976; Schlesinger, 1997), so only a small fraction of th e organic matter pool is dissolved at any time; (2) DOM-Pb complexes solubilized in upper soil hor izons may be precipitated or adsorbed lower in the soil profile; (3) the DOM to which Pb is bound may be utilized by microbes, allowing the associated Pb to bind anew to soil organic matter. Together, these factors tend to moderate the

release of Pb to surface waters in temperate terrestrial ecosyst ems. However, stressors or
disturbances that result in increased release of DOM from soils could result in the unanticipated release of Pb to groundwater and/or surface waters.

Acidification The effect of acidification on ecosystem cycling of Pb is difficult to predict. Like most metals, the solubility of Pb increases as pH decreases (Stumm and Morgan, 1995), suggesting that enhanced mobility of Pb should be found in ecosystems under acidification stress. However, Pb is also strongly bound to organic matter in soil s and sediments. Reductions in pH may cause a decrease in the solubility of DOM, due to the protonation of carboxylic functional groups (Tipping and Woof, 1990). Because of the importance of Pb complexation with organic matter,

lower DOM concentrations in soil solution result ing from acidification may offset the increased solubility of Pb and hence decrease the mobility of the organically bound metal. In a study of grassland and forest soils at th e Rothamsted Experiment Station in England, long-term (i.e., >100 years) soil acidification signif icantly increased the mob ility of Pb in the soilAX7-91 (Blake and Goulding, 2002). However, the increas ed mobility was only observed in very acid soils, those with pH of <4.5. The fraction of exchangeable Pb (extracted with 0.1 M CaCl 2 )
increased from about 3% to 15% of the total Pb in the most acidified soils. Similarly, the fraction of organically bound Pb increased from about 2% of total Pb in neutral soils to 12% of

total Pb in the most acidified soils. Simila rly, Nouri and Reddy (1995) obs erved higher levels of diethylenetriaminepentaacetic-acid-[DTPA]-extractab le Pb in soils in a loblolly pine forest treated with simulated acid rain, but only in the mo st acidic treatment, with simulated rain with a pH of 3.5. Although acidification may increas e the mobility of Pb in some soils, it is not clear that this Pb is actually moving through or out of the soil profile. In an examination of running waters in Sweden, Johansson et al. (1995) found no relationship between acidification and Pb concentrations in drainage wate rs and concluded that Pb conc entrations were governed by the DOM concentration, which masked any associ ation with acidification. In an in situ lysimeter study, Bergkvist (1986) measured lower concentr ations of Pb in soil solutions draining experimentally acidified plots th an in unacidified plots. In a laboratory study using large soil columns, Merino and García-Rodeja (1997) observe d no effect of experimental acidification on the release of Pb to soil solution. Thus, while ac idification may increase th e potential mobility of Pb in soils, as indicated by increases in labile soil fractions such as exchangeable and DTPAextractable Pb, the actual movement of Pb in the soil is limited by DOM solubilization and transport. It is worth noting that in all of these studies, significant effects of acidification were observed for other trace metals (Bergkvist, 1986 ; Johansson et al., 1995; Merino and GarcíaRodeja, 1997). Acidification may enhance Pb export to draina ge water in very sandy soils, soils with limited ability to retain organic matter. Studies in the McDonald’s Branch watershed in the New Jersey pine barrens, where soil texture is similar to beach sands , suggested little Pb retention in the mineral soil (Swanson and Johns on, 1980; Turner et al., 1985). If acidification results in the mobilization of Pb and organic ma tter into these mineral soils, then increased streamwater Pb concentrations would likely follow.AX7-92 Land Use and Industry Changes in land use also represent potentia lly significant changes in the cycling of organic matter in terrestrial ecosystems. C onversion of pasture and croplands to woodlands changes the nature and quantity of organic matter i nputs to the soil. In te mperate climates, forest ecosystems tend to accumulate organic matter in an O horizon on the forest floor, whereas organic matter in grasslands and agricultural fields is concentrated in an A horizon at the soil surface. Andersen et al. (2002) compared the tr ace metal concentrations in arable fields in Denmark to nearby sites that had been converted to forest land. After 34 years of afforestation, the soils showed no significant difference in Pb concentration or fracti onation, despite significant

acidification of the soils. Afforestation ha d no effect on the soil carbon concentration, suggesting that land use change may have little effect on Pb cycling unless soil carbon pools are

affected. Similarly, the introduction of industrial ac tivity may have consequences for organic matter cycling, and subsequently, Pb mobilization. In a rare long-term study of polluted soils, Egli et al. (1999) studied the changes in trace metal concentrations in forest soils at a site in

western Switzerland between 1969 and 1993. The site is 3 to 6 km downwind from an aluminum industrial plant that operated between the 1950s and 1991. In the 24-year period of study, the site experienced significant d eclines in organic carbon in su rface (0 to 5 cm depth) and subsurface (30 to 35 cm) soils. In the 30 to 35 cm layer, the organic carbon concentration declined by more than 75%. Extractable Pb (using an ammonium acetat e and EDTA mixture) declined by 35% in the same layer. The authors suggested that the Pb lost from the soil had been organically bound. While this study indicates that loss of soil carbon can induce the mobilization and loss of Pb from terrestrial ecosystems, it is also worth noting that the decline in soil Pb was considerably smaller than the decline in organi c carbon. This suggests that Pb mobilized during organic matter decomposition can resorb to rema ining organic matter or perhaps to alternate binding sites (e.g., Fe and Mn oxides). The effects of industries that emit Pb to the atmosphere are discussed in Sections

AX7.1.5.2 and AX7.1.5.3 below. Forest harvesting represents a severe disrupt ion of the organic matter cycle in forest ecosystems. Litter inputs are severely reduced for several years after cutting (e.g., Hughes and Fahey, 1994). The removal of the forest canopy resu lts in reduced interception of precipitation,AX7-93 and, therefore, increased water flux to the soil surface. Also, until a new canopy closes, the soil surface is exposed to increased solar radiation and higher temper atures. Together, the higher moisture and temperature in surface soils te nd to increase the rate of organic matter decomposition. Several studies have estimated decreases of up to 40% in the organic matter

content of forest floor soils af ter clear-cutting (Covington, 1981; Federer, 1984; Johnson et al., 1995b). This loss of organic matter from the fore st floor could result in the mobilization of organically complexed Pb. However, observations from clear-cut sites in the United States and Europe indicate that forest harvesting causes little or no mobilization of Pb from forest soils. At the Hubbard Brook Experimental Forest in New Hampshire, whole-tree harvesting, the most intensive form of clear-cutting, resulted in ve ry small increases in Pb concentrations in soil solutions draining the Oa soil hori zon despite substantia l reductions in the organic matter mass of that horizon (Fuller et al., 1988; Johnson et al., 1995b). These in creases were associated with similarly small increases in dissolved organi c carbon (DOC) concentratio ns in the Oa horizon soil water. Output of Pb from the watershed stream was unaffected by clear-cutting. Similarly, Berthelsen and Steinnes (1995) observed small d ecreases in the Pb content of the Oa horizon (“H” in the European system of soil classificatio n) in clear-cut sites in Norway, compared to uncut reference sites. This mobilization of Pb from the Oa horizon was accompanied by an increase in the Pb content of the upper minera l soil horizons. The Pb decline in the Oa horizon was accompanied by a decrease in the organic matter content, leading the au thors to attribute the Pb dynamics to leaching with DOM. In a st udy conducted in Wales, Durand et al. (1994) observed lower Pb outputs from a stream draining a clear-cut watershed than from where the stream drained the upper reaches of the wa tershed, which were uncut. The DOC and H + outputs were also lower in the clear-cut area. These pa tterns persisted in all 5 years of the study. Forest harvesting is a severe form of ecosyst em disturbance, and, t hus, it is somewhat surprising that studies of clear-cutting have show n little or no effect on Pb mobility or loss from forest ecosystems. Perhaps the strong complexa tion behavior of Pb with natural organic matter

results in the retention of Pb in forest soils. Ev en in cases where Pb is mobilized in forest floor soils (Fuller et al., 1988; Berthels en and Steinnes, 1995), there is no evidence of loss of Pb from the ecosystem, indicating that mineral soils are effi cient in capturing and re taining any Pb that is mobilized in the forest floor. Therefore, the princi pal risk associated with forest harvesting is the loss of Pb in particulate form to drainage wa ters through erosion. In some relatively remoteAX7-94 lakes in the United Kingdom th e distribution of sediment-bound trace elements (including Pb) have been affected by forestry activities a nd catchment erosion (Yang and Rose, 2005). Yang and Rose (2005) believe that more contaminated soil in-wash could increase sediment heavy

metal concentrations while less contaminated so il in-wash could dilute sediment heavy metal

levels.

Climate Change Atmospheric Pb is not likely to contribute si gnificantly to global climate change. Lead compounds have relatively short residence times in the atmosphere, making it unlikely that they will reach the stratosphere. Als o, Pb compounds are not known to absorb infrared radiation and, therefore, are unlikely to cont ribute to stratospheric ozone de pletion or global warming. Climate change does, however, represent a disturbance to terrestrial ecosystems. Unfortunately, the potential linka ges between climate-related stre ss and Pb cycling are poorly understood. As in the previous examples, effects related to alterations in organic matter cycling may influence Pb migration. For example, an in crease in temperature leading to increased rates of organic matter decomposition could lead to te mporary increases in DOM concentrations and smaller steady-state pools of soil organic matter. Either of these factors could result in increased

concentrations of Pb in waters draining terrestrial ecosystems. Climate change may also affect the fluctuati ons of temperature and/ or precipitation in terrestrial ecosystems. For exampl e, there is some evidence for recent increases in the frequency of soil freezing events in the nor theastern United States (Mitch ell et al., 1996). Soil freezing

occurs when soils have little or no snow cover to insulate them from cold temperatures and

results in an increased release of nitrate and DOC from the O hor izons of forest soils (Mitchell et al., 1996; Fitzhugh et al., 2001). Increased DOC losses from O horizons subjected to freezing may also increase Pb mobilization. Increased fluctuations in precipitation may induce more frequent flooding, with potentially significant consequences for Pb c ontamination of floodplain ecosystems. Soils collected from the floodplain of the Elbe River, in Germany, contained elevated concentrations of Pb and other trace metals (Kr η ger and Grongroft, 2003). Tissues of plants from floodplain sites did not, however, contain higher Pb concentr ations than control si tes. More frequentAX7-95 or more severe flooding would lik ely result in increased inputs of Pb and other metals to floodplain soils.

AX7.1.4.2 Effects of Lead Exposure on Natu ral Ecosystem Structure and Function The effects of Pb exposure on natural ecosy stems are confounded by the fact that Pb exposure cannot be decoupled fr om other factors that may also affect the ecosystem under consideration. Principal among these factors are other trace metals and acidic deposition. Emissions of Pb from smelting and other indus trial activities are acco mpanied by other trace metals (e.g., Zn, Cu, and Cd) and sulfur dioxide (SO 2 ) that may cause toxic effects independently

or in concert with Pb. Reductions in the use of alkyl-Pb additives in ga soline have resulted in significant decreases in Pb depos ition to natural ecosystems in the northeastern United States (Johnson et al., 1995a). However, the period in which Pb deposition has declined (ca. 1975 to the present) has also seen signi ficant reductions in the acidity (i .e., increased pH) of precipitation in the region (Likens et al., 1996; Driscoll et al., 1998). Theref ore, changes in ecosystem Pb fluxes may be the result of reduced Pb inputs and/or reduced acidity. Experimental manipulation studies do not su ffer from these confounding effects, because Pb can be added in specific amounts, with or without other co mpounds. Unfortunately, ecosystem-level manipulations involving Pb additi ons have not been undertaken. Therefore, we

must use observations from field st udies of Pb behavior in sites exposed to various forms of Pb pollution to assess the effects of Pb on terrestria l ecosystems. This section includes a discussion

of effects of Pb in the structure and function of terrestrial ecosystems. Effects on energy flows (food chain effects) and biogeochemical cy cling are discussed in Section AX7.1.5.3.

Sites Affected by Nearby Point Sources of Lead Natural terrestrial ecosystems near smelters , mines, and other industrial plants have exhibited a variety of effects re lated to ecosystem structure and function. These effects include decreases in species diversity, changes in floral and faunal community composition, and decreasing vigor of terrestrial vegetation.

All of these effects were observed in eco systems surrounding the Anaconda smelter in southwestern Montana, which operated between 1884 and 1980 (Galbraith et al., 1995). Soils in

affected areas around the Anaconda smelter were en riched in Pb, arsenic, copper, cadmium, andAX7-96 zinc; had very low pH; and were determined to be phytotoxic to native vegetation (Kapustka et al., 1995). The elevated soil arsenic and metal concentrations occurred despite significantly lower organic matter concentrations in affected so ils relative to reference sites (Galbraith et al., 1995). Line-transect measurements indicated that affected sites had an average of 6.9 species per 10-m of transect, compared to 20.3 species per 10- m in the reference areas. More than 60% of the reference sites supported coniferous (58 %) or deciduous (3%) forest communities, whereas
less than 1% of the affected si tes retained functioning forest st ands. Abundant dead timber and stumps confirmed that the affected sites were once as forested as the reference sites. Affected grassland sites were also less diverse and had higher abundances of invasive species than reference grasslands. More than 50% of the aff ected sites were classified as bare ground. The occurrence of bare ground was si gnificantly correlated with the phytotoxicity scores derived by Kapustka et al. (1995), indicating a link between phytotoxicity and th e loss of vegetation in the

affected area. Because of the plant community changes near the Anaconda smelter, the vertical diversity of habitats in the affected ecosystems decrease d, with only shrubs and soil remaining as viable habitats. Galbraith et al. (1995) also used the Bureau of Land Management’s habitat evaluation procedure (HEP) to estimate habitat suitability indices (HSI) for two indicator species, marten ( Martes americana ) and elk ( Cervus elaphus ). The HSI value ranges from 0 (poor habitat) to 1 (ideal habitat). In sites affected by the Anac onda smelter, HSI values for marten averaged 0.0, compared to 0.5 to 0.8 for the reference sites. For elk, affected sites had an average HSI of 0.10, compared to 0.31 at reference sites.

Similar observations were made in the ar ea surrounding Palmerton, Pennsylvania, where two zinc smelters operated between 1898 and 1980. Soils in the area were enriched in Cd, Zn, Pb, and Cu, with concentrations decreasing with distance from the smelter sites (Beyer et al., 1985; Storm et al., 1994). Smelting was determined to be the principal source of Pb in soils in residential and undeveloped areas around Palmert on (Ketterer et al., 2001), which lies on the north side of a gap in Blue Mountain, a ri dge running roughly east-w est in east-central Pennsylvania. Much of the north-facing side of Blue Mountain within 3 km of the town is bare ground or sparsely vegetated, whereas the surr ounding natural landscape is predominantly oak forest (Sopper, 1989; Storm et al., 1994). Biodivers ity in affected areas is considerably lower than at reference sites, a pattern attributed to emissions from th e smelters (Beyer et al., 1985;AX7-97 Sopper, 1989). The history is complicated, how ever, by the land use history of the area. Logging and fire in the early 20th century may al so have played a role in the changes in the terrestrial ecosystems (Jordan, 1975). Extens ive logging occurred after the smelters began operation, suggesting that some of the logging may have been salvage logging in affected areas. Regardless, the smelter emissions appear to have inhibited the regrowth of ecosystems compared to those in nearby unaffected areas. As in Anaconda, MT, the changes in the structure and function of the Palmerton ecosystem changed its suitability as a habitat for fauna that would normally inhabit the area. Storm et al. (1994) did not find amphibians or common invertebrates in two study sites nearest to the smelters. In the larger study area, they documented elevated concentrations of Pb, Cd, Cu, and Zn in tissues of species ranging in size from red-backed salamanders ( Pletheron cenereus ) to white-tailed deer ( Odocoilius virginianus ). Metal pollution around a Pb-Zn smelter near Bris tol, England has not resulted in the loss of oak woodlands within 3 km of the smelter, despite significant accumulation of Pb, Cd, Cu, and Zn in soils and vegetation (Martin a nd Bullock, 1994). However, the high metal concentrations have favored the growth of metal-tolerant species in the woodland.

The effects of Pb and other chemical emissi ons on terrestrial ecosystems near smelters

and other industrial sites decrea se downwind from the source. Several studies using the soil

burden as an indicator have shown that much of the contamination occurs within a radius of 20 to 50 km around the emission source (Miller and McFee, 1983; Martin and Bullock, 1994; Galbraith et al., 1995; Spurge on and Hopkin, 1996a; see also S ection AX7.1.2). For example,

the concentration of Pb in fore st litter declined downwind from a Pb-Zn smelter near Bristol, UK, from 2330 to 3050 ppm in a stand 2.9 km from the smelter to 45 to 110 ppm in a stand 23 km from the smelter (Martin and Bullock, 1994). Thus, while sites near point sources of Pb may experience profound effects on ecosystem struct ure and function, the extent of those effects is limited spatially. Elevated metal concentrations around smelte rs have been found to persist despite significant reduct ions in emissions (Hrsak et al., 2000). Most terrestrial ecosystems are far enough from point sources that long-range Pb transport is the primary mechanism for Pb inputs.AX7-98 Sites Affected by Long-Range Lead Transport Because the effects of anthropogenic Pb emi ssions tend to be restricted in geographic extent, most natural terrestrial ecosystems in the U.S. sites have Pb burdens derived primarily from long-range atmospheric trans port. Pollutant Pb represents a large fraction of the Pb in many of these ecosystems. In particular, many of these sites have accumulated large amounts of Pb in soils. For example, at the Hubbard Br ook Experimental Forest in New Hampshire, the

amount of Pb in the forest floor was estim ated to have increased from about 1.35 kg ha ! 1 in 1926 (before the introduction of alkyl-Pb additives in gasoline) to 10.5 kg ha ! 1 in 1977 (Johnson et al., 1995a). They also estimated the atmospheri c Pb deposition from 1926 to 1987 to be 8.7 kg ha ! 1

, an amount that could account for nearly all of the incr ease in Pb in the fo rest floor during the period. The input of precipitation Pb to th e Hubbard Brook ecosystem in the six decades spanning 1926 to 1987 was more than half of the total Pb estimated to have been released by mineral weathering in the entire 12,000- to 14,000-year post-glacial period (14.1 kg ha ! 1 : [Johnson et al., 2004]). Other studies employi ng Pb budgets (Miller and Friedland, 1994; Watmough et al., 2004), and Pb isotopes (Baco n et al., 1995, 1996; Watmough et al., 1998; Bindler et al., 1999; Hansmann and Köppel, 2000; Kaste et al., 2003), have also shown that pollutant Pb, primarily from gasoline combustion, re presents a quantitativel y significant fraction of labile Pb in temperate soils, especially in the upper, orga nic-rich horizons. Despite years of elevated atmospheric Pb inputs and elevated concentrations in soils, there is little evidence that sites affected primar ily by long-range Pb transport have experienced significant effects on ecosystem structure or f unction. Low concentrations of Pb in soil solutions, the result of strong complexation of Pb by soil organic matter, may explain why few ecological effects have been observed. At Hubba rd Brook, for example, the concentration of Pb in soil solutions draining the Oa horizon is < 0.1 μM and is even lower in solutions draining mineral-soil horizons (Driscoll et al., 1988; Wang et al., 1995). Friedland and Johnson (1985)

measured similar concentrations in soil solution s collected from deciduous and spruce-fir stands on Camel’s Hump Mountain in Vermont. In an undeveloped, forested watershed in Maryland, Scudlark et al. (2005) found that atmospheric input of some elements (Al, Cd, Ni, Zn) is effectively transmitted through the watershed, wher eas other elements (Pb, As, Se, Fe, Cr, Cu) are strongly sequestered, in the respective order noted.AX7-99 In ecosystems where Pb concentrations in so il solutions are low, toxicity levels for vegetation are not likely to be reached regardless of the soil Pb concentration. Furthermore, mycorrhizal infection of tree root s appears to reduce the translocat ion of Pb from roots to shoots (Marschner et al., 1996; Jentschke et al., 1998). In a study of mycorrhizal and non-mycorrhizal Norway spruce ( Picea abies (L.) Karst.), mycorrhizal infecti on of roots was not affected by Pb dose. Some, but not all, species of myco rrhizae showed reductions in the amount of extrametrical mycelium with Pb exposure but only at solution concentrations of 5 μM, a level at least 50 times greater than typical concentrations in forest soils. In a related study, the growth rate of mycorrhizal fungi was unaffected at so lution Pb concentrations of 1 and 10 μM, but decreased at 500 μM (Marschner et al., 1999).

Low soil solution Pb concentrations and the influence of mycorrhizal symbionts also result in low uptake of Pb by terrestrial vegetati on. The net flux of Pb into vegetation in the northern hardwood forest at Hubbard Brook in the 1980s was estimated as only 1 g ha ! 1 year ! 1 (Johnson et al., 1995a), representing 3% of the precipitation input. Klaminder et al. (2005) also measured a Pb uptake of 1 g ha ! 1 year ! 1 in a spruce-pine forest in northern Sweden. Despite plant uptake fluxes being very low, they are sensitive to differences and changes in Pb deposition. Berthelsen et al. (1995) observed decreases in the Pb content of stem, twig, leaf, and needle tissues of a variety of tree species in No rway between 1982 and 1992, when atmospheric Pb deposition declined by approximately 70%. Th ey also observed significantly lower Pb

concentrations in tree tissues collected in nor thern Norway versus southern Norway, where atmospheric Pb deposition is greater. Even at subtoxic concentrations, Pb and othe r metals may influence species diversity in terrestrial ecosystems. However, little work ha s been done on the effect of low-level metal concentrations on species diversity. In one study, plant species diversity was positively

correlated to the concentration of available Pb in natural and artificial urban meadows in Britain (McCrea et al., 2004). The aut hors hypothesized that Pb may inhibit phosphorous uptake by dominant species, allowing less abundant (but more Pb-tol erant) ones to succeed.

AX7.1.4.3 Effects of Lead on Energy Flows and Biogeochemical Cycling In terrestrial ecosystems, energy flow is clos ely linked to the carbon cycle. The principal input of energy to terrestrial ecosyste ms is through photosynthesis, in which CO 2 is converted toAX7-100 biomass carbon. Because of this link between phot osynthesis and energy flow, any effect that Pb

has on the structure and function of terrestrial ecosystems (as discu ssed in Section AX7.1.5.3.) influences the flow of energy into the ecosystem . This section focuses on how Pb influences energy transfer within terrestrial ecosystems, which begin with the de composition of litter and other detrital material by soil bacteria and fungi, and cascade th rough the various components of the detrital food web. Because the mobility of Pb in soils is closely tied to organic matter cycling, decomposition processes ar e central to the biogeochemical cycle of Pb. This section concludes with a discussion of how biogeochemical cycling of Pb has changed in response to the changing Pb inputs to terrestrial ecosystems.

Effects of Lead on Detrital Energy Flows Lead can have a significant effect on energy fl ows in terrestrial ecosyst ems. At some sites severely affected by metal pollu tion, death of vegetation can occu r, dramatically reducing the input of carbon to the ecosystem (Jordan, 1975; Galbraith et al., 1995). Subsequently, wind and erosion may remove litter and humus, leaving bare mineral soil, a nearly sterile environment in which very little energy transfer can take place (Little and Martin, 1972; Ga lbraith et al., 1995).

At Pb-affected sites that can re tain a functioning forest stand, the rate of decomposition of litter may be reduced, resulting in greater accumu lation of litter on the forest floor than in unpolluted stands. Numerous investigators ha ve documented significan t declines in litter decomposition rates (Cotrufo et al., 1995; Johns on and Hale, 2004) and/or the rate of carbon respiration (Laskowski et al., 1994; Cotrufo et al ., 1995; Saviozzi et al., 1997; Niklínska et al., 1998; Palmborg et al., 1998; Aka and Darici, 2004) in acid- and metal-contam inated soils or soils treated with Pb. The resulting accumulation of organic matter on the soil surface can be dramatic. For example, an oak woodland 3 km fr om a smelter in Bristo l, England had a litter layer mass 10 times greater than the mass in a si milar stand 23 km from the smelter (Martin and Bullock, 1994). Reduced decomposition rates in polluted ecosystem s are the result of the inhibition of soil bacteria and fungi and its effects on microbial community structure (Bååth, 1989). Kuperman and Carreiro (1997) observed 60% lower substrat e-induced respiration in heavily polluted grassland soils near th e U.S. Army’s Aberdeen Proving Grou nd in Maryland. This decline in carbon respiration was associated with 81% lo wer bacterial biomass and 93% lower fungalAX7-101 biomass. Similar declines in the activities of carbon-, nitrogen-, and phosphorus-acquiring enzymes were also observed. Such dramatic effects have only been observed in highly contaminated ecosystems. In a less contaminated grassland site near a Pb factory in Germany, Chander et al. (2001) observed a lower ratio of microbial biomass carbon to soil organic carbon in polluted soils. The ratio of basal respiration to microbial biomass (the “metabolic quotient,” qCO 2 ) declined with increasing metal concentr ation, though this observation depended on the procedure for measuring microbial biomass (subs trate-induced respiration versus fumigationextraction). The combined effect of lower micr obial biomass per unit soil carbon and similar or lower qCO 2 on polluted sites indicates that the ability of soil microorganisms to process carbon inputs is compromised by metal pollution. The type of ecosystem also plays a role in de termining the effects of Pb and other metals on the microbial processing of litter. Forest so ils in temperate zones accumulate organic matter at the soil surface to a greater degree than in gr asslands. This organic- rich O horizon can support a large microbial biomass; but it is also an effect ive trap for Pb inputs, because of the association between Pb and soil organic matter. At highly c ontaminated forest sites, microbial biomass and enzyme activities may be depressed (Fritze et al., 1989; Bååth et al., 1991), causing slower decomposition of the litter. In addition to effects on decomposition and carbon transformations, Pb and other trace metals can also influence key nitrogen cycling pr ocesses. Studies in th e 1970s demonstrated that

Pb and other metals inhibit the mineralizati on of nitrogen from soil organic matter and nitrification (Liang and Tabatabai, 1977, 1978), resulting in lower nitr ogen availability to plants. More recent research has documented significant inhi bitory effects of Pb and other metals on the activities of several enzymes believed to be crucial to nitrogen mineralization in soils (Senwo and Tabatabai, 1999; Acosta-Martinez and Tabataba i, 2000; Ekenler and Tabatabai, 2002). This suggests that the inhibitory effect of Pb and other metals is broa d-based, and not specific to any particular metabolic pathway. In reducing enviro nments, the rate of denitrification is also depressed by trace metals. Fu and Tabatabai (1989) found that 2.5 μmol g ! 1 of Pb (ca. 500 mg/kg ! 1 ) was sufficient to cause 0, 27, and 52% decr eases in nitrogen reductase activity in

three different soils. Metal pollution can also affect soil invertebrate populations. Martin and Bullock (1994) observed lower abundances of a variety of woodlice, millipedes, spiders, insects, and earthwormsAX7-102 in an oak woodland site 3 km from a Pb-Zn smelte r in Bristol, England, co mpared to a reference

site 23 km from the smelter. The differences were most dramatic when expressed per unit mass of litter. Several species that were abundant in the reference site were not found in the contaminated woodland. For exampl e, the abundance of the woodlice Trichoniscus pusillus was 151 individuals per m 2 in the reference woodland, but none were found in the contaminated

soils. This was also true of 2 of the 3 m illipede species, and 4 of the 5 earthworm species

studied. At six sites within 1 km from the sm elters, no earthworms were present at all (Spurgeon and Hopkin, 1996a). Contamination at this site has apparently reduced both the population and biodiversity of the soil in vertebrate community. The effect of metal pollution on soil invertebrates may be a threshold-type response. In a study conducted in woodlands near two zinc sme lters in Noyelles-Godault, in northern France, soils at the most polluted site were devoid of mites and millipedes, while the remaining sites had diversity measures similar to cont rol sites (Grelle et al., 2000). While Pb pollution affects the population and di versity of soil fauna, there is little evidence of significant bioaccumulation of Pb in the soil food web (see also Section AX7.1.3). In the Bristol, England study, Pb concentratio ns in earthworms were lower than soil Pb

concentrations and much lower than litter Pb c oncentrations (Martin and Bullock, 1994). Litterdwelling mites had Pb concentrations that were 10% of the average litte r concentration. The predator centipedes Lithobius forficatus and L. variegatus had mean Pb concentrations of 18.6 and 44.0 mg kg ! 1 , respectively, two orders of magnitude lower than the Pb concentration of litter (2193 mg kg ! 1 ) and lower than the concentrations of their known prey species. In a study conducted in a Norway spruce forest affected primarily by automobile exhaust from a nearby highway, earthworms had Pb concentrations simila r to the soil (Roth, 1993) . Almost all of the litter decomposers, however, had Pb concentrations that were less than 20% of the litter. All but 3 of the zoophagous arthropods had Pb concentrations that were less than 40% of their prey; the remaining 3 had Pb concentrations similar to thei r prey. Because of the absence of significant bioaccumulation in the soil food web, predator speci es will be affected by Pb pollution primarily through effects on the abundance of thei r prey (Spurgeon and Hopkin, 1996b). Taken as a whole, ecosystem-level studies of the soil food web indicate that Pb can affect energy flows in terrestrial ecosystems through two principal mechanisms. In the most severely polluted sites, the death of primary producers di rectly decreases the flow of energy into theAX7-103 ecosystems. More commonly, the accumulation of toxi c levels of Pb or other metals in litter and soil decreases the rate of litter decomposition th rough decreases in microbial biomass and/or respiration. These reductions can subsequently a ffect higher trophic levels that depend on these organisms. It is important to note that sites that have exhibited signi ficant disruption to energy

flows and the terrestrial food web are sites that have experienced severe metal contamination and

adverse effects from SO 2 from smelters or other metals-related activities.

Lead Dynamics in Terrestrial Ecosystems Lead inputs to terrestrial ecosystems in the Unite d States have declined dramatically in the past 30 years, primarily because of the almost complete elimination of alkyl-Pb additives in gasoline in North America. Also, Pb emissions from smelters have declined as older plants have been shut down or fitted with improved emissi ons controls. Unfortunately, there are few longterm data sets of precipitation inputs to te rrestrial ecosystems. At the Hubbard Brook Experimental Forest, in New Hampshire, Pb in put in bulk deposition declined by more than 97% between 1976 and 1989 (Johnson et al., 1995a). Studies of freshwater sediments also indicate a dramatic decline in Pb inputs since the mid- 1970s (Graney et al., 1995; Johnson et al., 1995a; Farmer et al., 1997; Br ännvall et al., 2001a,b). Reported concentrations of Pb in waters dr aining natural terrestri al ecosystems have always been low (Wang et al., 1995; Bacon and Bain, 1995; Johnson et al., 1995a; Vinogradoff et al., 2005), genera lly less than 1 ng L ! 1 , even at moderately pollu ted sites (Laskowski et al., 1995). Consequently, most terrestrial ecosystems in North America and Europe remain sinks for Pb despite reductions in atmospheric Pb deposi tion of more than 95%. At Hubbard Brook, for example, the input of Pb in bulk precipitation declined from 325 g ha ! 1 year ! 1 between 1975 and 1977 compared to 29 g ha ! 1 year ! 1 between 1985 and 1987 (Johnson et al., 1995a). During the same period, the output of Pb in stream water declined from 6 g ha ! 1 year ! 1 to 4 g ha ! 1 year ! 1

. Thus, despite the decline in Pb input, 85% of the incoming Pb was still reta ined in the terrestrial ecosystem in the later time period. Similar obser vations have been made in Europe, where the use of leaded gasoline has also declined in th e last few decades. At the Glensaugh Research Station in Scotland, the input of Pb to the forest ecosystem was estimated as 42.6 g ha ! 1 year ! 1
between 2001 and 2003, about six time s the stream export of 7.2 g ha ! 1 year ! 1 (Vinogradoff et al., 2005). Similarly, Huang and Matzner (2004) reported a throughfall flux of 16.5 g ha ! 1 year ! 1 atAX7-104 the forested Lehstenbach catchment in Bavaria, about six times the efflux in runoff of 2.82 g ha ! 1

year ! 1

. Lead pollution has resulted in the accumulation of large Pb burdens in terrestrial ecosystems (see also Section AX7.1.2). Despite reduc tions in emissions, this accumulation of Pb continues, though at markedly lower rates. Th e large pool of Pb bound in soils may potentially be a threat to aquatic ecosystems (see Section AX7.2), depending on its rate of release from the soil. Early estimates of the residence time of Pb in the forest floor ranged from 220 to 5,000 years (Benninger et al., 1975; Friedland and J ohnson, 1985; Turner et al., 1985). However, more recent literature suggests that Pb is transported mo re rapidly within soil profiles than previously believed. The pool of Pb in forest floor soils of the northeastern United States declined significantly in the late 20th century. Friedland et al. (1992) reported a 12% decline in the

amount of Pb in forest floor soils at 30 site s in the region between 1980 and 1990, a much greater decline than would be expected for a pool with a residence time of 220 to 5,000 years. At Hubbard Brook, the pool of Pb in the fo rest floor declined by 29% between 1977 and 1987, an even more rapid rate of loss than reported by Friedland et al. (1992) . More recently, Evans et al. (2005) reported significan t declines in the Pb content of forest floor soils in the

northeastern United States and eastern Canada between 1979 and 1996. The magnitude of the decrease in Pb content was greates t at their sites in southern Verm ont, and smallest at sites on the Gaspe Peninsula in Quebec, reflecting the historic gradient in Pb deposition in the region. At the Vermont site, the Pb concentrati on in the litter layer (Oi horizon) was 85% lower in 1996 than in 1979. In the Gaspe peninsula of Quebec, the decrease was only 50%. Since drainage water Pb concentrations remain low, the Pb released from forest floor soils in the past has been largely immobilized in mineral soils (Miller a nd Friedland, 1994; Johnson et al., 1995a; Johnson and Petras, 1998; Watmou gh and Hutchinson, 2004; Johnson et al., 2004). This is supported by evidence from Pb-isot ope analyses. Gasoline-derived Pb has a 206 Pb: 207 Pb ratio that can be easily discriminated from Pb in the rocks from which soils are derived. Using isotopic mixing models with gasoline-Pb and Pb in soil parent materials as end members,

a number of researchers have documented the ac cumulation of pollutant Pb in mineral soils (Bindler et al., 1999; Kaste et al., 2003; Watmough and Hutchinson, 2004; Bacon and Hewitt, 2005; Steinnes and Friedland, 2005). In a ha rdwood stand on Camel’s Hump Mountain in Vermont, as much as 65% of the pollutant Pb deposited to the stand ha d moved into mineralAX7-105 horizons by 2001 (Kaste et al., 2003). In a spruce-fir stand, cont aining a thicker organic forest floor layer, penetration of pollutant Pb into the mineral soil was much lower.

This recent research has resulted in a reeval uation of the turnover time of Pb in forest floor soils. The Camel’s Hump data suggest that Pb resides in the forest floor of deciduous stands for about 60 years and about 150 years in coniferous stands (Kaste et al., 2003). These values are somewhat greater than those publis hed previously by Miller and Friedland (1994), who used a Pb budget approach. Extremely ra pid turnover of Pb was observed in some

hardwood forest floor soils in south-central Ontario (Watmough et al., 2004). Their estimated turnover times of 1.8 to 3.1 years are much lowe r than any other published values, which they attribute to the mull-type forest floor at their sites. Mull-type forest floors are normally underlain by organic-rich A horizons, capable of immobilizing Pb released from the forest floor. Indeed, at the same site in Ontario, Watmough and Hutchi nson (2004) found that 90% of the pollutant Pb

could be found in this A horizon. The time period over which the accumulated Pb in soils may be released to drainage waters remains unclear. If Pb moves as a pulse through the soil, there may be a point in the future at which problematic Pb concentrations occur. However, several authors have argued against this hypothesis (Wang and Benoit, 1997; Kaste et al., 2003; Watmough et al., 2004), contending that the strong linkage between Pb and DOM will result in a temporally dispersed release of Pb in the form of Pb-DOM complexes. Thus, the greatest threat is likely to be in the most highly contaminated areas surrounding point sources of Pb, where the amount of Pb accumulated in the soil is high, and the death of vegetation has resulted in reduced soil organic

matter levels.

AX7.1.4.4 Summary Atmospheric Pb pollution has resulted in the accumulation of Pb in terrestrial ecosystems throughout the world. In the Unite d States, pollutant Pb represen ts a significant fraction of the total Pb burden in soils, even in sites remote from smelters and other industrial plants. However, few significant effects of Pb pollution have been obs erved at sites that are not near point sources of Pb. Evidence from precipitation collection and sediment analyses indicates that atmospheric deposition of Pb has declined dramatically (>95%) at sites unaffected by point sources of Pb, andAX7-106 there is little evidence that Pb accumulated in soils at these sites represents a threat to groundwaters or surface water supplies. The highest environmental risk for Pb in terrest rial ecosystems exists at sites within about 50 km of smelters and other Pb-emitting indus trial sites. Assessing the risks specifically associated with Pb is difficult, because these si tes also experience elevated concentrations of

other metals and because of effects related to SO 2 emissions. The concentrations of Pb in soils, vegetation, and fauna at these sites can be two to three orders of magnitude higher than in reference areas (see Sections AX7.1.2 and AX7.1.4.2). In the most extreme cases, near smelter sites, the death of vegetation causes a near-compl ete collapse of the detr ital food web, creating a terrestrial ecosystem in which energy and nutrien t flows are minimal. More commonly, stress in soil microorganisms and detritivores can cause reducti ons in the rate of decomposition of detrital organic matter. Although there is little evidence of significant bioa ccumulation of Pb in natural terrestrial ecosystems, reductions in microbial and detritivorous populations can affect the success of their predators. Thus , at present, industrial point so urces represent the greatest Pbrelated threat to the maintenance of sustainable, healthy, diverse, and hi gh-functioning terrestrial ecosystems in the United States.



### AX7.2 AQUATIC ECOSYSTEMS AX7.2.1 Methodologies Used in Aquatic Ecosystem Research As discussed in previous sections, aerial deposition is one source of Pb deposition to aquatic systems. Consequently, to develop air quality criteria for Pb, consideration must be given to not only the environmental fate of Pb, but also to the environmental effects of Pb in the aquatic environment through consid eration of laboratory toxicity studies and field evaluations. Perhaps the most straightforward approach for eval uating the effects of Pb is to consider extant criteria for Pb in aquatic ecosystems, i.e., wate r and sediment quality criteria. A key issue in developing Pb water and sediment criteria that are broadly appli cable to a range of water bodies is properly accounting for Pb bioa vailability and the range in speci es sensitivities. This section

summarizes how these criteria are derived, the types of toxicity studies considered, and key factors that influence the bioavailability of Pb in surface water and sediment to aquatic life. Because Pb in the aquatic environment is often associated with other metals (e.g., cadmium,AX7-107 copper, zinc), the importance of considering the toxicity of metal mixtures is also discussed. Finally, some issues related to background Pb con centrations are briefly addressed. It is beyond the scope of this section to review all methodologies in aqua tic system research, but good reviews can be found in summary books , such as Rand et al. (1995).

AX7.2.1.1 Analytical Methods Common analytical methods for measuring Pb in the aquatic environment are summarized

in Table AX7-2.1.1. For relevance to the ambien t water quality criteria (AWQC) and sediment quality criteria for Pb discussed below, minimum detection limits should be in the low parts per billion (ppb) range for surface water and the lo w parts per million (ppm) range for sediment.

Table AX7-2.1.1. Common Analytical Methods for Measuring Lead in Water, Sediment, and Tissue Analysis Type Analytical Method Direct-Aspiration (Flame) Atomic Absorption Spectroscopy (AAS) EPA SW-846 Method 7420 a

, EPA Method 239.1 b

, Standard Method 3111 c Graphite Furnace Atomic Absorption Spectroscopy (GFAAS) EPA SW-846 Method 7421 a

, EPA Method 239.2 b

, Standard Method 3113 c Inductively Coupled Plasma (ICP) EPA SW-846 Method 6010B a

, EPA Method 200.7 b

, Standard Method 3120 c Inductively Coupled Plasma-Mass Spectrometry (ICP-MS) EPA SW-846 Method 6020 a

, EPA Method 200.8 b

a U.S. Environmental Protection Agency (1986c) Test Methods for Evaluating Solid Waste, Physical/Chemical

Methods (SW-846). Third Edition, September 1986; Final Updates I (7/92), IIA (8/93), II (9/94), IIB (1/95), III (12/96), IIIA (4/98), IIIB (11/04). b U.S. Environmental Protection Agency (1991) Methods for the Determination of Metals in Environmental

Samples. EPA/600/4-91-010. June 1991 (Supplement I, EPA/600/R-94-111, May 1994).

c American Public Health Association (1995) Standard Methods for the Examination of Water and Wastewater, 19th Edition. American Public Health Association, American Water Works Association, Water Pollution Control Federation.

In addition to the methods presented in Table AX7-2.1.1, many of the methods discussed in Section AX7.1.1 can be applied to suspended solids and sediments collected from aquaticAX7-108 ecosystems. Just as in the terrestrial environmen t, the speciation of Pb and other trace metals in natural freshwaters and seawater plays a crucia l role in determining their reactivity, mobility, bioavailability, and toxicity. Many of the same speciation techniques employed for the speciation of Pb in terrestrial ecosystems (s ee Section AX7.1.1 and AX7.1.2) are applicable in aquatic ecosystems.

There is now a better understa nding of the potential effects of sampling, sample handling, and sample preparation on aqueous-phase metal sp eciation. Thus, a need has arisen for dynamic analytical techniques that are able to capture a metal’s speciation, in -situ and in real time. Some of these recently developed dynamic trac e metal speciation techniques include:

• Diffusion gradients in thin-film gels (DGT)

• Gel integrated microelectrodes combined with voltammetric in situ profiling (GIME-VIP)

• Stripping chronopoten tiometry (SCP)

• Flow-through and hollow fiber permeation liquid membranes (FTPLM and HFPLM)

• Donnan membrane technique (DMT)

• Competitive ligand-exchange/stripping voltammetry (CLE-SV)

Various dynamic speciation techniques were co mpared in a study by Si gg et al. (2006) using

freshwaters collected in Switzer land. They found that techni ques involving in-situ measurement (GIME-VIP) or in-situ exposure (DGT, DMT, and HFPLM) appeared to the most appropriate for avoiding Pb and other trace metal speciation artif acts associated sampling and sample handling.

AX7.2.1.2 Ambient Water Qualit y Criteria: Development The EPA’s procedures for deriving AWQC are described in Stephan et al. (1985) and are summarized here. With few exceptions, AWQC are derived based on data from aquatic toxicity

studies conducted in the laborator y. In general, both acute (sho rt term) and chronic (long term) AWQC are developed. Depending on the species, th e toxicity studies considered for developing acute criteria range in length from 48 to 96 hours. Acceptable endpoints for acute AWQC development are mortality and/or immobilization, expressed as the median lethal concentration (LC 50 ) or median effect concentration (EC 50 ). For each species, the geometric mean of the acceptable LC 50 /EC 50 data is calculated to determine th e species mean acute value (SMAV).AX7-109 For each genera, the geometric mean of the releva nt SMAVs is then calculated to determine the genus mean acute value (GMAV). The GMAVs are then ranked from high to low, and the final acute value (FAV; the 5th per centile of the GMAVs, based on the four GMAVs surrounding the 5th percentile) is determined. Because the FAV is based on LC 50 /EC 50 values (which represent unacceptably high levels of effect), the FAV is di vided by two to estimate a low-effect level.

This value is then termed the acute criterion, or crite rion maximum concentration (CMC). Based on the most recent AWQC document for Pb (U .S. Environmental Protection Agency, 1985),

Table AX7-2.1.2 shows the freshwater SMAVs and GMAVs for Pb, and the resulting freshwater CMC. Note that the freshwater AWQC are norma lized for the hardness of the site water, as

discussed further below in Section AX7.2.1.3.

Table AX7-2.1.2. Development of Current Acute Freshwater Criteria for Lead (U.S. Environmental Protection Agency, 1985) 1

Rank Species GMAV (μg/L) SMAV (μg/L) 10 Midge ( Tanytarsus dissimilis ) 235,900 235,900

9 Goldfish ( Carassius auratus ) 101,100 101,100

8 Guppy ( Poecilia reticulata ) 66,140 66,140

7 Bluegill ( Lepomis macrochirus ) 52,310 52,310

6 Fathead minnow ( Pimephales promelas ) 25,440 25,440

5 Brook trout ( Salvelinus fontinalis ) 4,820 4,820

4 Rainbow trout ( Oncorhynchus mykiss ) 2,448 2,448

3 Snail ( Aplexa hypnorum ) 1,040 1,040

2 Cladoceran ( Daphnia magna ) 447.8 447.8

1 Amphipod ( Gammarus pseudolimnaeus ) 142.6 142.6

FAV = 67.54 μg/L

CMC = 33.77 μ g/L

1 All values are normalized to a hardness of 50 mg/L (see Section AX7.2.1.3).AX7-110 To develop chronic AWQC, acceptable chronic toxicity studies should encompass the full life cycle of the test organism, although for fish, early life stage or partia l life cycle toxicity studies are considered accepta ble. Acceptable endpoints include reproduction, growth and development, and survival, with the effect levels expressed as the chronic value, which is the geometric mean of the no-obser ved-effect concentration (NOEC) 1 and the lowest-observedeffect concentration (LOEC) 2 . Although a chronic criterion c ould be calculated as the 5th percentile of genus mean chronic values (GMCVs), sufficient chronic toxic ity data are generally lacking, as is the case for Pb. Consequently, an acute-chronic ratio (ACR) is typically applied to

the FAV to derive the chronic criterion. As the name implies, the ACR is the ratio of the acute

LC 50 to the chronic value, based on studies with th e same species and in the same dilution water. For Pb, the final ACR is 51.29, which results in a final chronic value (FCV) of 1.317 μg/L (at a hardness of 50 mg/L). The U.S. EPA guidelines fo r developing AWQC (Stephan et al., 1985) are now more than 20 years old and thus are not re flective of scientific advances in aquatic toxicology and risk assessment that have developed since the 1980s. For example, the toxicological importance of dietar y metals has been increasingly recognized and approaches for incorporating dietary metals into regulatory criteria are being ev aluated (Meyer et al., 2005).

Other issues include consideration of certain su blethal endpoints that are currently not directly incorporated into AWQC development (e.g., endo crine toxicity, behavi oral responses) and protection of threatened and e ndangered (T&E) species (U.S. E nvironmental Protection Agency, 2003). In deriving appropriate and scientifically defensible air qua lity criteria for Pb, it will be important that the state-of-the-science for metals toxicity in aquatic systems be incorporated into the development process. Subsequent sections summari ze some of the toxicity studies that meet the AWQC development guidelines, with an emphasis on key studies published since the last Pb AWQC

were derived in 1984.

1 The NOEC is the highest concentration tested that di d not result in statistically significant effects relative to the control.

2 The LOEC is the lowest concentration tested that re sulted in statistically significant effects relative to the control.AX7-111 AX7.2.1.3 Ambient Water Quality Cr iteria: Bioavailability Issues In surface waters, the environmental fate of metal contaminants is mitigated through adsorption, complexation, chelation, and other proce sses that affect bioavaila bility. The toxicity

of divalent cations tends to be hi ghest in soft waters with low c oncentrations of dissolved organic matter and suspended particles. In an acidic envi ronment (pH <4), the ioni c form of most metals generally predominates and is considered to be the more toxic form. As the pH increases, carbonate, oxide, hydroxide, and sulfide complexes of the metals tend to predominate, and tend to be less toxic (Flore nce, 1977; Miller and Mackay, 1980). The portion of dissolved metal available for uptake or bioaccumulation is influe nced by modifying factors that “sequester” the metal in an environmental matrix, thereby reducing the bioavailability of the metal at the sites of action. Metals can become comp lexed (bound) to a ligand that can make metals either more toxic (via transport mechanisms) or less toxic (by changing the metal’s biological activity). Metals that complex tightly to ligands generally are not readil y bioavailable and, thus, are less toxic to aquatic biota than th eir free-metal ion counterpart s (Carlson et al., 1986; McCarthy, 1989). There are many kinds of ligands, organic and inorganic, as well as natural and manmade. Ligands found in natural surface waters a nd municipal and industrial effluent discharges include glycine, ammonia, oxalate, humic or fulvic acids, hydroxide , carbonate, bicarbonate, chloride, and hydrogen sulfide (Stumm and Morg an, 1970; Martin, 1986; Pagenkopf, 1986). Recognizing the importance of calcium and magnesium ions (hardness) in modifying Pb toxicity, the current freshwater AWQC for Pb are normalized based on the hardness of the site water. The acute freshwater criteria, for exampl e, are 30, 65, and 136 μg/L at hardness levels of 50, 100, and 200 mg/L (as CaCO 3 ). Although it has been known for some time that other water quality parameters such as pH, dissolved or ganic carbon (DOC), and alkalinity affect the bioavailability of metals to aqua tic biota, it was the relatively recent development of the biotic ligand model (BLM) that allows for AWQC to pote ntially consider all of these factors. Paquin et al. (2002) provided a thorough re view of the factors influenci ng metal bioavailability and how

research over the last few decades has culmin ated in the development of the BLM. By understanding the binding affinities of vari ous natural ligands in surface waters and how the freshwater fish gill inte racts with free cations in the water, one can predict how metals exert their toxic effects (Schwart z et al., 2004). Models developed prior to the BLM are the freeion activity model (FIAM) and the gill surface interaction model (GSIM). The FIAM accountsAX7-112 for the binding of free-metal ion and other metal complexes to the site of toxic action in an organism; and it also considers competition be tween metal species and other cations (Paquin et al., 2002). The GSIM is fundamentally simila r to the FIAM in that it accounts for competition between metal ions and hardness cations at the ph ysiological active gill sites, but whereas the FIAM is largely conceptual, the GSIM was used in interpreting toxicity test results for individual metals and metal mixtures (Pagenkopf, 1983). The BLM was adapted from the GSIM and uses the biotic ligand, rather than the fish gill as the site of toxic action (Di Toro et al., 2001; Paquin et al., 2002). This approach, ther efore, considers that the extern al fish gill surface contains receptor sites for metal binding (Schwartz et al., 2004) and that acute toxicity is associated with the binding of metals to defined sites (biotic ligands) on or with in the organism (Paquin et al., 2002). The model is predicated on the theory that mortality (or other toxic effects) occurs when the concentration of metal bound to biotic ligan d exceeds a threshold concentration (Di Toro et al., 2001; Paquin et al., 2002). Free metal catio ns “out compete” other cations and bind to the limited number of active receptor sites on the gill surface, which may ultimately result in suffocation and/or disruption of io noregulatory mechanisms in the fi sh, leading to death (Di Toro et al., 2001; Paquin et al., 2002). Because the BLM uses the biotic ligand (not the fish gill) as the site of action, the model can be applied to other aquatic orga nisms, such as crustaceans, where the site of action is directly exposed to the aqueous environment (D i Toro et al., 2001). Although the BLM is currently be ing considered as a tool fo r regulating metals on a sitespecific basis, there are potential limitations in using the BLM to regulate metals in surface

waters that should be understood in developing air quali ty criteria for lead. For example, BLMs developed to-date have focused on acute mort ality/immobilization endpoints for fish and invertebrates. Chronic exposures are typically of greatest regulatory concern, but chronic BLMs to date have received limited attention (D e Schamphelaere and Janssen, 2004). In addition, BLMs account for uptake of dissolved metal, but dietary metals have also been shown to contribute to uptake by aquatic biot a and, in some cases, increased toxicity. Besser et al. (2005) observed that chronic (42-day) Pb toxicity to the amphipod Hyalella az teca was greater from a combined aqueous and dietary exposure than from a water-only exposure and the authors conclude that estimates of chronic toxicity th resholds for Pb should consider both aqueous and dietary exposure routes. The feasibility of incorporating dietary metals into BLMs is under investigation. Another important issue that mu st be addressed in developing and applying aAX7-113 BLM in AWQC development is the type of disso lved organic matter used in model development relative to the types of dissolved organic matter at the site of interest. Richar ds et al. (2001) demonstrated that natural organic matter (NOM) of different types differentially influenced Pb and Cu accumulation by gills of rainbow trout. Th ere are also other ligands not accounted for in the BLM that require more research. Bianch ini and Bowles (2002) emphasized the importance of reduced sulfur as a metal ligand, limitations in scientific knowledge on reduced sulfur, and provided recommendations for studi es necessary to incorporate sulfide ligands into the BLM. The U.S. EPA is currently revising the aquatic life AWQC for lead, which will include toxicity data published after th e 1985 AWQC were released and incorporation of the BLM is being evaluated. If an acute BLM is incorporat ed into the revised AWQC for lead, the chronic criteria would likely be estimated using an A CR (the same approach used when the acute criterion is based on empirical toxicity data; see above).

AX7.2.1.4 Sediment Quality Criteria: Development and Bioavailability Issues As with metals in surface waters, the environmental fate of metal contaminants in sediments is moderated through various binding pro cesses that reduce the concentration of free,

bioavailable metal. Sediments function as a si nk for Pb, as with most metals. Lead compounds such as Pb-carbonates, Pb-sulfates, and Pb-sul fides predominate in se diments (Prosi, 1989). Total Pb has a higher retention time and a higher per centage is retained in sediments compared to copper and zinc (Prosi, 1989). Lead is primar ily accumulated in sediments as insoluble Pb complexes adsorbed to suspended particulate matter. Naturally occurring Pb is bound in sediments and has a low geochemical mobility (Prosi, 1989). Organic-sulfide and moderately reducible fractions are less mobile, whereas catio n-exchangeable fractions and easily-reducible fractions are more mobile and more readily bi oavailable to biota (Prosi, 1989). Most Pb transported in surface waters is in a particulate fo rm, originating from the erosion of sediments in rivers or produced in the wa ter column (Prosi, 1989). Sediment quality criteria have yet to be adopted by the EPA, but an equilibrium partitioning procedure has recently been publis hed (U.S. Environmental Protection Agency, 2005c). The EPA has selected an equilibrium partitioning approach because it explicitly accounts for the bioavailability of metals. This approach is based on mixtures of cadmium, copper, Pb, nickel, silver, and zinc. Equilibrium partitioning (EqP) theory predicts that metalsAX7-114 partition in sediment between aci d-volatile sulfide, pore water, benthic organisms, and other sediment phases such as organic carbon. When the sum of the molar concentrations of simultaneously extracted metal ( ∑ SEM) minus the molar concentration of AVS is less than zero, it can accurately be predicted that sediments are not toxic because of these me tals. Note that this approach can be used to predict the lack of t oxicity, but not the presen ce of toxicity. It is important to emphasize that metals must be ev aluated as a mixture us ing this approach. If individual metals, or just two or th ree metals, are measured in sediment, ∑ SEM would be misleadingly small and it may inaccurately appear that ∑ SEM / AVS is less than 1.0.

If ∑ SEM / AVS is normalized to the organic carbon fraction (i.e., ( ∑ SEM / AVS)/ f

OC ), mortality can be more reliably predicted by acco unting for both the site-specific organic carbon and AVS concentrations. When evaluating a me tal mixture containing cadmium, copper, Pb, nickel, silver, and zinc, the following predicti ons can be made (U.S. Environmental Protection Agency, 2005c):

• A sediment with (SEM / AVS)/ f

OC < 130 μmol/g OC should pose low risk of adverse biological effects due to these metals.

• A sediment with 130 μmol/g OC < (SEM / AVS)/ f

OC < 3000 μmol/g OC may have adverse biological effects due to these metals.

• In a sediment with (SEM / AVS)/ f

OC > 3000 μmol/g OC , adverse biological effects may be expected.
A third approach is to measure pore water c oncentrations of cadmium, copper, Pb, nickel,

and zinc and then divide the concentrations by their respective FCVs. If the sum of these quotients is <1.0, these metals are not expected to be toxic to benthic organisms. It should be noted that although EPA ha s endorsed the AVS-SEM approach for developing sediment guidelines, ot her investigators do not necessa rily agree, as the AVS-SEM approach may not be relevant to benthic organisms that ingest sedi ment particles. For example, Griscom et al. (2002) found that metals associat ed with either reduced or oxidized sediment particles can be assimilated by deposit and suspen sion feeding bivalve species due to the low pH and moderate reducing conditions in bivalve guts. Although Lee et al. (2000) agreed that the AVS-based approach may be appropriate for pr otecting some benthic organisms from acute toxicity associated w ith exposure to very high pore water concentrations of metals, they questioned whether the AVS-based approach is relevant to less contaminated sediments underAX7-115 more natural conditions with ver tical stratification of oxygen concentrations and, hence, varying levels of AVS. The authors concluded that impor tant uncertainties remain in the application of the AVS-SEM approach as a regulatory tool. Similarly, some studies suggest that AVS-SEM measurements in the natural environment must be interpreted cautiously as AVS can be quite variable with sediment depth and season (Van de n Berg et al., 1998). Finally, Long et al. (1998) critically compared the ability of AVS-SEM and dry weight-normalized con centrations to predict toxicity of sediment-associated tr ace metals. Based on an analysis of 77 field-collected sediment samples, Long et al. (1998) found that dry weig ht-normalized concentrations were equally or slightly more accurate than AVS-SEM in predic ting non-toxic and toxic results in laboratory bioassays. Long et al. suggests th at a limitation of the AVS-SEM a pproach is that the criteria

were not derived from field studie s with mixture of toxic chemical s and thus may be less relevant to sites comprising complex chemical mixtures . Thus, although the AVS-SEM approach for developing sediment quality criter ia is being pursued by the U.S. EPA, there is clearly not scientific consensus on this approach, at least not for all circumstances. Many alternative approaches for developing sediment quality guidelines are based on empirical correlations between metal concentrations in sediment to associated biological effects, based on sediment toxicity test s (Long et al., 1995; Ingersoll et al., 1996; MacDonald et al., 2000). However, these guidelines are based on tota l metal concentrations in sediment and do not account for the bioavailability of metals betw een sediments. Sediment quality guidelines proposed for Pb from these other sources are shown in Table AX7-2.1.3.

Table AX7-2.1.3. Recommended Sediment Quality Guidelines for Lead

Source Water Type Guideline Ty pe Conc. (mg/kg dw) MacDonald et al. (2000) Freshwater TEC PEC 35.8 128 Ingersoll et al. (1996) Freshwater ERL ERM 55 99 Long et al. (1995) Saltwater ERL ERM 46.7 218

TEC = Threshold effect concentration; PEC = Probable effect concentration; ERL = Effects range – low; ERM = Effects range – medianAX7-116 AX7.2.1.5 Metal Mixtures As discussed above, the EPA’s current approach for developing sediment criteria for Pb and other metals is to consider the molar sum of the metal concentrations ( ∑ SEM). Although a similar approach has not been a pplied to AWQC, metal mixtures have been shown to be more toxic than individual metals (Spe har and Fiandt, 1986; Enserink et al., 1991). Spehar and Fiandt (1986) evaluated the acute and chronic toxicity of a metal mixture (arsenic, cadmium, chromium, copper, mercury, and Pb) to fathead minnows ( Pimephales promelas ) and a daphnid ( Ceriodaphnia dubia ). In acute tests, the joint toxicity of these metals was observed to be more

than additive for fathead minnows and nearly strictly additive for daphnids. In chronic tests, the joint toxicity of the metals was less than a dditive for fathead minnows and nearly strictly additive for daphnids. One approach for considering the additive toxi city of Pb with other metals is to use the concept of toxic uni ts (TUs). Toxic units for each component of a metal mixture are derived by dividing metal concentrations by their respective acute or chro nic criterion. The TUs

for all the metals in the mixture are then summed. A ∑ TU > 1.0 suggests the metal mixture is toxic (note that this is the sa me approach as discussed above for developing metal sediment criteria based on pore water concentrations). According to Norwood et al. (2003), the TU approach is presently the most appropriate mode l for predicting effects of metal mixtures based on the currently available toxicity data. Howeve r, it should also be emphasized that the TU approach is most appropriate at a screening level, because the tr ue toxicity of the mixture is dependent on the relative amounts of each metal. The TU approach is also recommended with mixtures containing less than six metals.

Lead and other metals often co-occur in sedi ments with other toxicants, such as organic contaminants. Effects-based sediment quality gu idelines (SQGs) have been developed over the past 20 years to aid in the in terpretation of the relationships between complex chemical contamination and adverse biological effects (Long et al., 2006). Mean sediment quality guideline quotients (mSQGQs) can be calculated by dividing the concentrations of chemicals in sediments by their respective SQGs and then ca lculating the mean of the quotients for the individual chemicals. Long et al. (2006) perfor med a critical review of this approach and found that it reasonably predicts the incidence and magnitude of toxici ty in laboratory tests and the incidence of impairment to benthic commun ities increases incrementally with increasing mSQGQs. However, the authors po inted out some of the limitations of this approach, such as aAX7-117 lack of agreement on the level of mSQGQs, masking of an indivi dual chemical’s effect due to data aggregation, lack of SQGs for all chemi cals of concern, and mSQGQs were not initially derived as a regulatory standard or criterion, thus there is a reluctance to use them in

enforcement or remediati on (Long et al., 2006). For assessing Pb effects on aquatic ecosystems, it is not truly feasible to account for metal mixtures, because these will obviously vary highly fr om site to site. However, the toxicity of

metal mixtures in surface water should be considered on a site-specific basis.

AX7.2.1.6 Background Lead Because Pb is naturally occurring, it is found in all environmental compartments including surface water, sediment, and aquatic biot a. Background Pb concentrations are spatially variable depending on geological fe atures and local characteristic s that influence Pb speciation and mobility. In the European Union risk assessments for metals, an “added risk” approach has been considered that assumes only the amount of metal added above background is relevant in a toxicological evaluation. However, this a pproach ignores the po ssible contribution of background metal levels to toxic effects, and background metal levels are regionally variable, precluding the approach from be ing easily transferable between sites. In terms of deriving environmental criteria for Pb, b ackground levels should be consider ed on a site-specific basis if

there is sufficient information that Pb concentr ations are naturally el evated. As discussed previously, the use of radiogenic Pb isotope s is useful for sour ce apportionment.



### AX7.2.2 Distribution of Lead in Aquatic Ecosystems Atmospheric Pb is delivered to aquatic ecosystems primarily through deposition (wet and/or dry) or through erosiona l transport of soil pa rticles (Baier and Healy, 1977; Dolske and Sievering, 1979; and Yang and Rose, 2005). A numb er of physical and chemical factors govern

the fate and behavior of Pb in aquatic systems. The EPA summa rized some of these controlling factors in the 1986 Lead AQCD (U .S. Environmental Protection Agency, 1986a). For example, the predominant form of Pb in the environment is in the divalent (Pb 2+ ) form and complexation with inorganic and organic li gands is dependent on pH (L overing, 1976; Rickard and Nriagu, 1978). A significant portion of Pb in the aquatic environment exists in the undissolved form (i.e., bound to suspended particulate ma tter). The ratio of Pb in susp ended solids to Pb in filtrateAX7-118 varies from 4:1 in rural streams to 27:1 in urban streams (Getz et al ., 1977). In still waters, Pb is removed through sedimentation at a rate determ ined by temperature, pH, oxidation-reduction (redox) potential, organic conten t, grain size, and chemical form of Pb in the water and biological activities (J enne and Luoma, 1977). Since the publication of the 1986 Lead AQCD (U.S. Environmental Protection Agency, 1986a), know ledge of the properties of Pb in aquatic ecosystems has expanded. This section will provid e further detail on the chemical species and

the environmental factors affecti ng speciation of Pb in the aqua tic environment. In addition, quantitative distributions of Pb in water, sedime nt, and biological tissues will be presented for aquatic ecosystems throughout the United States. Finally, recent studies discussing the tracing of Pb in aquatic systems will be summarized.

AX7.2.2.1 Speciation of Lead in Aquatic Ecosystems The speciation of Pb in the aquatic envir onment is controlled by many factors. The primary form of Pb in aquatic environments is divalent (Pb 2+ ), while Pb 4+ exists only under extreme oxidizing conditions (Rickard and Nriagu, 1978). Labile forms of Pb (e.g., Pb 2+

, PbOH + , PbCO 3 ) are a significant portion of the Pb inputs to aquatic systems from atmospheric washout. Lead is typically present in acidic aquatic environments as PbSO 4 , PbCl 4 , ionic Pb, cationic forms of Pb-hydroxide, and ordinary Pb -hydroxide (Pb(OH) 2 ). In alkaline waters, common species of Pb include ani onic forms of Pb-carbonate (Pb(CO 3 )) and Pb(OH) 2 . Speciation models have been developed based on the chemical equilibrium model developed by Tipping (1994) to assist in examining metal speciation. The EPA MINT EQA2 computer model (http://www.epa.gov/ceampubl/mmedia/minteq/) is one such equilibrium speciation model that

can be used to calculate the e quilibrium composition of dilute a queous solutions in the laboratory or in natural aqueous systems. The model is useful for calculating the equilibrium mass distribution among dissolved species, adsorbed species, and multiple solid phases under a variety of conditions, including a gas phase with constant partial pressures. In addition to chemical equilibrium models, the speciation of metals is important from a toxicological perspective. The BLM was developed to study the toxicity of metal ions in a quatic biota and was previously

described in Section AX7.2.1.3. Fu rther detail on speciation m odels is not provided herein, rather a general overview of major speciation principles are characterized in the following

sections.AX7-119 Acidity (pH)

Freshwater Most of the Pb in aquatic environments is in the inorganic form (Sadiq, 1992). The speciation of inorganic Pb in freshwater aquatic ecosystem s is dependent upon pH and the available complexing ligands. Solubility vari es according to pH, temperature, and water hardness (Weber, 1993). Lead ra pidly loses solubility above pH 6.5 (Rickard and Nriagu, 1978)

and as water hardness increases. In freshwater s, Pb typically forms strong complexes with inorganic OH ! and CO 3 2 ! and weak complexes with Cl ! (Long and Angino, 1977; Bodek et al., 1988). The primary form of Pb at low pH ( # 6.5) is predominantly Pb 2+ and less abundant inorganic forms include Pb(HCO) 3 , Pb(SO4) 2 2 ! , PbCl, PbCO 3 , and Pb 2 (OH) 2 CO 3 (Figure AX7-2.2.1). At higher pH ( ∃ 7.5), Pb forms hydroxide complexes (PbOH + , Pb(OH) 2 , Pb(OH) 3 ! , Pb(OH) 4 2 ! ).

Figure AX7-2.2.1. Distribution of aqueous lead species as a function of pH based on a concentration of 1 μg Pb/L (U.S. Environmental Protection Agency, 1999).AX7-120 Organic compounds in surface waters may originate from natural (e.g., humic or fulvic acids) or anthropogenic sources (e .g., nitrilotriacetoni trile and ethylenediaminetetraacetic acid [EDTA]) (U.S. Environmental Protection Agenc y, 1986b). The presence of organic complexes

has been shown to increase the rate of solu tion of Pb bound as Pb-sulfide (Lovering, 1976). Soluble organic Pb compounds are present at pH values near 7 and may remain bound at pH values as low as 3 (Lovering, 1976 ; Guy and Chakrabarti, 1976). At higher pH (7.4 to 9), Pb-organic complexes are partially decompose d. Water hardness and pH were found to be important in Pb-humic acid interactions (O ’Shea and Mancy, 1978). An increase in pH increased the concentration of exchangeable Pb complexes, while an increase in hardness tended

to decrease the humic acid-Pb interactions. Thus, the metals involved in water hardness apparently inhibit the exchangeable inte ractions between metals and humic acids.

Marine Water In marine systems, an increase in salin ity increases complexi ng with chloride and

carbonate ions and reduces the amount of free Pb 2+ . In seawaters and estu aries at low pH, Pb is primarily bound to chlorides (PbCl, PbCl 2 , PbCl 3 !

 , PbCl 4 2 ! ) and may also form inorganic Pb(HCO) 3 , Pb(SO4) 2 2 ! , or PbCO 3 . Elevated pH in saltwater environments results in the formation of Pb hydroxides (PbOH + , Pb(OH) 2 , Pb(OH) 3 ! , Pb(OH) 4 2 ! ) (Figure AX7-2.2.2). A recent examination of Pb species in seawater as a function of chloride concentration suggested that the primary species were PbCl 3 ! > PbCO 3 > PbCl 2 > PbCl + > and Pb(OH) + (Fernando, 1995).

Lead in freshwater and seawater systems is highly complexed w ith carbonate ligands suggesting that Pb is likely to be highly available fo r sorption to suspended materials (Long and Angino, 1977). Current information suggests that inorgani c Pb is the dominant form in seawater; however, it has been shown that organically bound Pb complexes ma ke up a large portion of the total Pb (Capodaglio et al., 1990).

Sorption Sorption processes (i.e., partitioning of dissolv ed Pb to suspended particulate matter or sediments) appear to exert a dominant effect on the distribution of Pb in the environment (U.S. Environmental Protection Agency, 1979). Sorp tion of Pb results in the enrichment ofAX7-121 Figure AX7-2.2.2. Lead speciation versus chloride content (Fernando, 1995).

bed sediments, particularly in environments with elevated organic matter content from anthropogenic sources. Lead adso rption to aquatic sediments is co rrelated with pollution in sites containing high levels of anthr opogenic organic content, even unde r acidic conditions (Tada and Suzuki, 1982; Brook and Moore, 1988; Davis an d Galloway, 1993; Botelho et al., 1994; Davis et al., 1996). Particulate-bound forms are more of ten linked to urban runo ff and mining effluents (Eisler, 2000). Solid Pb complexes form when Pb precipitate s or adsorbs to suspended particulate matter and sediments. Inorganic Pb adsorption to susp ended organic matter or sediments is dependent on parameters such as, pH, salinity, water hard ness, and the composition of the organic matter (U.S. Environmental Protection Agency, 1979). In addition to suspended organic matter, Pb can adsorb to biofilms (i.e., bacteria) (Nelson et al ., 1995; Wilson et al., 2001). Adsorption typically increases with increasing pH, increasing amounts of iron or manganese; and with a higher degree of polarity of the particulate matter (e.g., clay s). Adsorption decreases with water hardness (Syracuse Research Corporation, 1999). At higher pH, Pb precipitates as Pb(OH) + and PbHCO 3 + into bed sediments (Weber, 1993). Conversely, at low pH, Pb is negatively sorbed (repelled from the adsorbent surface) (U.S. Environmen tal Protection Agency, 1979; Gao et al., 2003). In addition, Pb may be remobilized from sediment with a decrease in metal concentration in theAX7-122 solution phase, complexation with chelating ag ents (e.g., EDTA), and changing redox conditions (Gao et al., 2003). Changes in water chemistr y (e.g., reduced pH or ionic composition) can

cause sediment Pb to become remobilized and potentially bioavailable to aquatic organisms (Weber, 1993).

Biotransformation Methylation may result in Pb remobiliza tion and reintroduction into the aqueous environment compartment and its subsequent rel ease into the atmosphere (Syracuse Research Corporation., 1999). However, methylation is not a significant e nvironmental pathway controlling the fate of Pb in the aquatic environmen t. The microbial methylation of Pb in aquatic systems has been demonstrated experimentall y, but evidence for natura l occurrence is limited (Beijer and Jernelov, 1984; DeJonghe and Adams, 1986). Reisinger et al. (1981) examined the methylation of Pb in the presence of numerous b acteria known to alkylate metals and did not find evidence of Pb methylation unde r any test condition. Tetramet hyl-Pb may be formed by the methylation of Pb-nitrate or Pb-chloride in sediments (Bode k et al., 1988). However, tetramethyl-Pb is unstable and may degrade in ae robic environments after being released from sediments (U.S. Environmental Protection Agenc y, 1986b). Methylated species of Pb may also be formed by the decomposition of tetralkyl -Pb compounds (Radojevic and Harrison, 1987; Rhue et al., 1992). Sadiq (1992) reviewed the me thylation of Pb compou nds and suggested that chemical methylation of Pb is the dominant pr ocess and that biomethylation is of secondary importance.

AX7.2.2.2 Spatial Distribution of Lead in Aquatic Ecosystems National Water Quality Assessment (NAWQA) The 1986 Lead AQCD (U.S. Environmental Pr otection Agency, 1986b) did not describe

the distribution and concentrati on of Pb throughout aquatic ecos ystems of the United States. Consequently, an analysis of readily availabl e data on Pb concentrations was conducted to

determine the distributio n of Pb in the aquatic environment. Data from the United States Geological Survey (USGS) National Water-Q uality Assessment (NAWQA) program were queried and retrieved. NAWQA contains data on Pb concentrations in surface water, bed sediment, and animal tissue for more than 50 ri ver basins and aquifers throughout the country,AX7-123 and it has been used by the EPA for describing na tional environmental concentrations for use in developing AWQC. The authors recognize that the NAWQA progr am encountered analytical challenges with the chemical analysis of Pb in surface waters. The analytical methods available during the NAWQA program were not as sensi tive as methods currently applied today. Therefore, analytical detection limits are elevat ed and a large portion of the data set contains non-detected values. Nevertheless, this data provides a comprehensive overview of Pb concentrations in U.S. surface waters that is suppl emented with data from other relevant studies. The following sections describe the estimated concentrations of Pb from NAWQA and other research programs. NAWQA data are collected duri ng long-term, cyclical invest igations wherein study units undergo intensive sampling for 3 to 4 years, followed by low-intensity monitoring and assessment of trends every 10 years. The NAW QA program’s first cycle was initiated in 1991; therefore, all available data are less than 15 years old. The second cycle began in 2001 and is ongoing; data are currently av ailable through 30 September 2003. The NAWQA program study units were selected to represent a wide variet y of environmental conditions and contaminant sources; therefore, agricultural, urban, and natura l areas were all included. Attention was also given to selecting sites covering a wide vari ety of hydrologic and ecological resources. NAWQA sampling protocols are designed to pr omote data consistency within and among study units while minimizing lo cal-scale spatial va riability. Water-column sampling is conducted via continuous monitoring, fixed-interval sampling, extreme-flow sampling, as well as seasonal, high-frequency sampling in order to characterize spatial, temporal, and seasonal variability as a function of hydr ologic conditions and contaminant sources. Sediment and tissue samples are collected during low-flow periods during the summer or fall to reduce seasonal variability. Where possible, sediment grab samples are collected al ong a 100-m stream reach, upstream of the location of the water-column samp ling. Five to ten deposit ional zones at various depths, covering left bank, right bank, and cente r channel, are sampled to ensure a robust representation of each site. Fine-g rained samples from the surficial 2 to 3 cm of bed sediment at each depositional zone are sampled and composited. Tissue samples are collected following a National Target Taxa list and d ecision trees that help guide selection from that list to accommodate local variability.AX7-124 The NAWQA dataset was chosen over other re adily available national databases (i.e. the USEPA-maintained database for the STOrage a nd RETrieval [STORET] of chemical, physical, and biological data), because the study design an d methods used to assess the water quality of each study unit are rigorous and consistent, and, as such, these data may be presented with a high

level of confidence. This is in stark contrast to the STORET database, which essentially serves as a depot for any organization wishing to share data they have generated. This lack of a consistent methodology or QA/QC pr otocol has lead to the STORET data being highly qualified and offered with only a mild leve l of confidence. Furthermore, because there is no standard for site selection within STORET, the database may be biased toward contaminated sites. Finally, and, perhaps most importantly, the majority of th e available Pb data in STORET predate the use of clean techniques for Pb quantification. The authors recognize the existen ce of several local and regiona l datasets that may be of quality equal to NAWQA; however, due to the nati onal scope of this assessment, these datasets were not included in the following statistical an alyses. However, because the NAWQA database

does not cover lakes and the marine/estuarine en vironment, and we were unable to identify any monitoring data of similar quality, local and regional datasets were used in these cases to provide general information on environmental Pb concentrations.

Data Acquisition and Analysis The following data were downloaded for the enti re United States (all states) from the NAWQA website (http://water.us gs.gov/nawqa/index.html): site information, dissolved Pb concentration in surface water (μ g/L), total Pb concentration (μg/g) in bed sediment (<63 μm) 3

, and Pb concentration in animal tissue (μg/g dw). Using the land use classification given for each site, the data were divided into two groups: “natural” and “ambient” (Table AX7-2.2.1). All samples were considered to fall within th e ambient group (the combined contribution of natural and anthropogenic sources), whereas the natural group comprised “forest,” “rangeland,” or “reference” samples only 4 . These groups follow those de fined and recommended for use by the EPA’s Framework for Inorganic Metals Risk Assessment (U.S. Environmental Protection

3 NAWQA sediment samples are sieved to <63 μm to promote the collection of fine-grained surficial sediments, which are natural accumulators of trace elements.

4 The authors acknowledge th at while Pb samples collected from site s classified under these three land use categories will most closely reflect natural background conc entrations, atmospheric input of lead may be present.AX7-125 Table AX7-2.2.1. NAWQA Land Use Categories and Natural/Ambient Classification

NAWQA Land Use Categories Classification Agricultural Ambient
Commercial/Industrial Ambient Cropland Ambient
Forest Ambient/Natural Mining Ambient
Mixed Ambient
NA Ambient Orchard/Vineyard Ambient
Other/Mixed Ambient
Pasture Ambient Rangeland Ambient/Natural
Reference Ambient/Natural
Residential Ambient
Urban Ambient

Agency, 2004b). Finally, in addition to the natu ral/ambient classification, tissue samples were further divided into “whole organism” and “liver” groups. All data were compiled in spreadsheets wher ein non-detect values we re converted to onehalf the detection limit and the total number of samp les, percentage of non- detect values (percent censorship), minimum, maximum, median, mea n, standard deviation, and cumulative density functions were calculated for each endpoint for both the natural and ambient groups. As discussed below, some datasets were highly censored; however, deletion of non-detect data

has been shown to increase the relative error in th e mean to a greater extent than inclusion of non-detects as 1⁄2 of the detection limit (Newma n et al., 1989); therefore means and other statistics were calculated using the latter method for this analysis . Finally, since all data were geo-referenced, a geographic information system (GIS; ArcGIS) was used to generate maps, conduct spatial queries and analys es, and calculate statistics.AX7-126 Lead Distributions Generate d from the NAWQA Database Natural versus Ambient Groups There were four to eight times more ambient surface water (Table AX7-2.2.2) and bulk sediment (Table AX7-2.2.3) samples in the compiled dataset than natural samples. This is most likely a function of both the NAWQA program site selection process and the fact that sites unaffected by human activities are extremely limite d. The spatial distributions of natural and ambient surface water/sediment sites were fairly comparable, with natural samples located in almost all of the same areas as ambient sample s except in the Midwest (Ohio, Illinois, Iowa, and Michigan), where natural site s were not present (Figure AX7 -2.2.3). This exception may be because these areas are dominated by agricu ltural and urban areas. The same spatial

distributions were observed for the natural and ambient liver and whole organism tissue samples (Figure AX7-2.2.4 and Figure AX7-2.2.5).

Table AX7-2.2.2. Summary Statistics of Ambient and Natural Levels of Dissolved Lead in Surface Water

Surface Water Dissolved Pb (μg/L)

Statistic Natural Ambient % Censorship 87.91 85.66

N 430 3445
Minimum 0.04 0.04
Maximum 8.40 29.78

Mean 0.52 0.66 Standard Deviation 0.59 1.20
Median 0.50 0.50
90th Percentile 0.50 0.50
95th Percentile 0.50 1.10
96th Percentile 0.67 2.00
97th Percentile 1.00 2.34
98th Percentile 1.79 3.58
99th Percentile 2.48 5.44AX7-127 Table AX7-2.2.3. Summary Statistics of Ambien t and Natural Levels of Total Lead in <63 μm Bulk Sediment

Bulk Sediment <63 um Total Lead (μg/g)

Statistic Natural Ambient % Censorship 1.16 0.48

N 258 1466
Minimum 0.50 0.50
Maximum 12000 12000

Mean 109.07 120.11
Standard Deviation 786.74 672.41
Median 22.00 28.00 90th percentile 66.30 120.00 95th percentile 161.50 200.00

Surface Water The total number of surface water Pb sample s was 3,445; however these data were highly censored with 85.66% of the ambient samples (2951/3445) and 87.91% of the natural samples (378/430) below the detection limit 5 (Table AX7-2.2.2). Consequently, the majority of the variability between these two datasets fell between the 95th and 100th (maximum) percentiles, as was shown by the frequency distributions of the two groups deviating only at the upper and lower tails with most of the overlapping data fa lling at 0.50 μg/L (one-hal f of the most common detection limit, 1.0 μg/L; Figure AX7-2.2.6). As expected, due to the definitions of the natural and ambient groups, the 95th and 100th percentile s were consistently higher for the ambient samples than the natural samples. Similarly, the mean ambient Pb concentration (0.66 μg/L) was higher than the mean natura l Pb concentration (0.52 μg/L). 6

5 The NAWQA dataset contains multiple detection limits for Pb in surface water that have decreased over time. While the majority of data were analyzed with a detection limit of 1.0 μg/L (before 2000/2001), the most recent samples were analyzed with either a 0.5, 0.2, 0.16, or 0.08 μg/L det ection limit (after 2000/2001), and some older samples (N = 20) were analyzed with a detection limit of 2.0 μg/L.

6 The same pattern was observed upon calculating the mean Pb concentra tions based on detect data only (ambient mean = 1.66 μg/L, natural mean = 0.87 μg/L); however, as previously discussed, calculations included non-detect data as 1⁄2 of the detection limit to reduce the relative error in the mean.AX7-128Figure AX7-2.2.3. Spatial distribution of natural and ambient surface water/sediment sites (Surface water: natural N = 430,ambient N = 3445; Sediment: natural N = 258, ambient N = 1466).AX7-129Figure AX7-2.2.4. Spatial distribution of natural and ambient liver tissue sample sites (Natural N = 83, Ambient N = 559).AX7-130Figure AX7-2.2.5. Spatial distribution ofnatural and ambient whole organism tissue sample sites (Natural N = 93,Ambient N = 332).AX7-131 Figure AX7-2.2.6. Frequency distribution of ambi ent and natural levels of surface water dissolved lead (μg/L).

Due to the preponderance of non-detectable (ND) measurements, assessing national trends

in surface water-dissolved Pb concentrations was not possible. However, areas with elevated Pb concentrations were identified by classifying the data with detect able Pb concentrations above and below the 99th percentile. The 99th percent ile (versus the 95th perc entile) was chosen in this instance to represen t extreme conditions given the small window of variability in the dataset. By convention, the 95th percentile wa s used in subsequent analyses of this type. Areas with high

surface water Pb concentrations were observed in Washington, Idaho, Utah, Colorado, Arkansas, and Missouri (Figure AX7-2.2.7). The maximum m easured Pb concentration was located in Canyon Creek at Woodland Park, ID, a site classified as mining land use.

Sediment There were approximately one-half of the number of surface water data available for sediments (N = 1466). In contrast to the surface water data, however, very few sediment data

were below the detection lim it (7/1466 ambient ND, 3/258 natural ND; Table AX7-2.2.3). As expected, the mean ambient Pb concentr ation was higher than the mean natural PbAX7-132Figure AX7-2.2.7. Spatial distribution of dissolved lead insurface water (N = 3445).`AX7-133 concentration (120.11 and 109.07 μg/g, respectivel y). Similarly, the median ambient Pb concentration was higher than the median natural Pb concentration (28.00 and 22.00 μg/g, respectively) and the ambient 95th percentil e was higher than the natural 95th percentile (200.00 and 161.50 μg/g, respectively). While th e natural and ambient surface water Pb distributions differed only at the extremes, the natu ral sediment Pb percentil es were consistently lower than the ambient percentiles throughout th e distributions (Figure AX7-2.2.8). Unlike the surface water dataset, because the sediment data set was not heavily censored, assessing national trends in sediment Pb concentrations was possibl e. The data were mapped and categorized into the four quartiles of the fre quency distribution (Figure AX7-2.2.9) . The following observations

were made:

• Sediment Pb concentrations generally increa sed from west to east (the majority of sites along East Coast had Pb concentrations in the fourth quartile of the sediment Pb concentration frequenc y distribution).

• Several “hot spots” of concen trated sites with elevated sediment Pb concentrations were apparent in various western states.

• Sediment Pb concentrations were genera lly lowest in the midwestern states (the majority of sites in North Dakota, Nebraska, Minnesota, and Iowa had Pb concentrations in the first or second quartile of sediment Pb concentration frequency distribution).

As was seen with surface water Pb concentr ations, the highest measured sediment Pb concentrations were found in Idaho, Utah, a nd Colorado. Not surpri singly, of the top 10 sediment Pb concentrations recorded, 7 were meas ured at sites classified as mining land use.

Tissue As was true for the surface water data, ther e were a high number of tissue samples below the detection limit (47/93 natu ral whole organism ND, 130/332 ambient whole organism ND, 74/83 natural liver ND, 398/559 ambient liver ND; Table AX7-2.2.4). In general, more non-censored data were available for whole orga nism samples than liver samples, and for ambient sites than natural sites. As expected, fo r whole organism samples, the 95th percentile Pb

concentration measured at ambien t sites was higher than that m easured at natural sites (3.24 and 2.50 μg/g, respectively); however, Pb liver conc entration 95th percen tiles for ambient and natural samples were very similar, with the natural 95th percentile actually higher than theAX7-134 Figure AX7-2.2.8. Frequency distribution of ambien t and natural levels of bulk sediment <63 μm total Pb (μg/g).

ambient 95th percentile (1.26 and 1.06 μg/g, respectiv ely). In addition, as expected, the median

and mean Pb liver concentrations of ambien t samples (0.15 and 0.36 μg/g, respectively) were higher than the median and mean Pb liver con centrations of natural samples (0.11 and 0.28 μg/g, respectively). The same pattern was observed in the whole organism median and mean Pb concentrations (ambient: me dian = 0.59, mean = 1.03; natural: median = 0.35, mean = 0.95 μg/g). In addition, the frequency distribut ions of the liver and whole organism Pb

concentrations followed the same trends, with the natural percenti les consistently lower than the ambient percentiles throughout the distribut ions (Figure AX7-2.2.10 and Figure AX7-2.2.11). These whole organism results were compared with findings from the 1984 U.S. Fish and Wildlife Service (USFWS) National Contaminan t Biomonitoring Program (NCBP) (Schmitt and Brumbaugh, 1990). As part of this program, 321 co mposite samples of 3 to 5 whole, adult fish of a single species were collected from 109 river and Great Lake stations throughout the country. Samples were analyzed for Pb concentrations (μg/g ww) and the geometric mean, maximum, and 85th percentile were calculated. Upon comparing these summary statistics with the equivalent NAWQA ambient group value (NCBP stations were representative of both natural and anthropogenically influenced conditions), a very strong agreement between the two analyses wasAX7-135Figure AX7-2.2.9. Spatial distribution of totallead in bulk sediment <63 μm (N = 1466).AX7-136 Table AX7-2.2.4. Summary Statistics of Ambient and Natural Levels of Lead in Whole Organism and Liver Tissues

Tissue Pb (μg/g dry weight)

Whole Organism Liver

Statistic Natural Ambient Natural Ambient

% Censorship 50.54 39.16 89.16 71.20

N 93 332 83 559

Minimum 0.08 0.08 0.01 0.01

Maximum 22.60 22.60 3.37 12.69

Mean 0.95 1.03 0.28 0.36

Standard Deviation 2.53 1.74 0.54 0.96

Median 0.35 0.59 0.11 0.15

90th percentile 1.40 2.27 0.37 0.59

95th percentile 2.50 3.24 1.26 1.06

observed for each endpoint (Table AX7-2.2.5). For example, NCBP and NAWQA geometric mean Pb concentrations were nearly identical (0.55 and 0.54 μg/g dw, respectively) and the 85th percentiles only differed by 0.5 μg Pb/g dw (NCBP, 1.10 and NAWQA, 1.60). The authors acknowledge that a high degree of censorship is present in both of these datasets and no firm conclusions can be drawn by comparing these means. The objective of this exercise was limited to showing how the NAWQA data co mpare to other national datasets. As was the case with surface water data, the high amount of non-detectable measurements

did not allow for a national assessm ent of spatial trends in Pb tissue concentrations. Instead, areas with high Pb tissue concentr ations were identified by classi fying the data above and below the 95th percentile. Similar to surface water and sediments, tissu e concentrations were found to be elevated in Washington, Idaho, Utah, Colora do, Arkansas, and Missouri; however, several of the highest measured Pb concentrations were al so found in study units in the southwestern andAX7-137 Figure AX7-2.2.10. Frequency distribution of ambi ent and natural levels of lead in liver tissue (μg/g dry weight).

Figure AX7-2.2.11. Frequency distribution of ambien t and natural levels of lead in whole organism tissue (μg/g dry weight).AX7-138 Table AX7-2.2.5. Comparison of NCBP and NAWQA Ambient Lead Levels in Whole Organism Tissues

Whole Organism Lead Concentration (μg/g dry weight)

Statistic NCBP 1 NAWQA

Geometric Mean 0.55 0.54

Maximum 24.40 22.60

85th Percentile 1.10 1.60

1 To convert between wet and dry weight, wet weig ht values were multiplied by a factor of five.

southeastern states (Figure AX7-2.2.12 and Figur e AX7-2.2.13). As expected, the majority of the samples with elevated Pb concentrations were taken from sites classified as urban, commercial/industrial, or mining.

Input and Distribution of Lead in Other Aquatic Systems Because the NAWQA database does not c over lakes or the sea where atmospheric deposition of Pb is highly likely, the primary litera ture was searched for studies using ultra-clean sampling/analytical techniques to characterize Pb concentrations in these environments. Lead concentrations in lakes and oceans were generally found to be much lower than those measured in the lotic waters assessed by NAWQA. Surface wa ter concentrations of dissolved Pb measured in Hall Lake, Washington in 1990 ranged from 2.1 – 1015.3 ng/L (Balistrieri et al., 1994). Nriagu et al. (1996) found that the average surfac e water dissolved Pb co ncentrations measured in the Great Lakes (Superior, Erie, and On tario) between 1991 and 1993 were 3.2, 6.0, and 9.9 ng/L, respectively. Lead concentrations ra nged from 3.2 – 11 ng/L acro ss all three lakes. Similarly, 101 surface water tota l Pb concentrations measured at the HOT station ALOHA between 1998 and 2002 ranged from 25 – 57 pmol/kg (5 – 11 ng/kg; (Boyle et al., 2005). Based on the fact that Pb is predominately found in the dissolved form in the open ocean (<90%; Schaule and Patterson, 1981), dissolved Pb concen trations measured at these locations would likely have been even lower than the total Pb concentrations reported.AX7-139Figure AX7-2.2.12. Spatial distribution oflead in liver tissues (N = 559).AX7-140Figure AX7-2.2.13. Spatial distribution of lead in wholeorganism tissues (N = 332).AX7-141 In open waters of the North Atlantic the declin e of Pb concentrations has been associated with the phasing out of leaded gasoline in No rth America and western Europe (Véron et al., 1998). Likewise, Pb restrictions in gasoline appear to have been effective in reducing atmospheric Pb loading to the Okefenokee Sw amp in southern Georgia/northern Florida (Jackson et al., 2004). Based on se diment cores from the Okefenokee Swamp, Pb concentrations were approximately 0.5 mg/kg prior to indus trial development, reached a maximum of approximately 31 mg/kg from about 1935 to 1965, a nd following passage of the Clean Air Act in 1970 concentrations have declined to about 18 mg /kg in 1990 (Jackson et al., 2004). Trends in metals concentrations (roughly 1970-2001) in sedime nt cores from 35 reservoirs and lakes in urban and reference settings were analyzed by Mahl er et al. (2006) to determine the effects of three decades of legislation, regulation, and chan ging demographics and industrial practices in

the United States on concentrations of metals in the environment. The researchers found that decreasing trends outnumbered increasing trends for all seven metals analyzed (Cd, Cr, Cu, Pb, Hg, Ni, and Zn). The most consistent trends we re for Pb and Cr: For Pb, 83% of the lakes had decreasing trends and 6% had in creasing trends; for Cr, 54% of the lakes had decreasing trends and none had increasing trends. Mass accumulati on rates of metals in cores, adjusted for background concentrations, decreased from th e 1970s to the 1990s, where median changes ranged from 246% (Pb) to 23% (Hg and Zn). Th e largest decreases were found in lakes located

in dense urban watersheds where the overall metals contamination in recently deposited

sediments decreased to on-half its 1970s median value. Howe ver, Mahler et al. (2006) found that anthropogenic mass accumulation rates in dens e urban lakes remained elevated over those in lakes in undeveloped watersheds, in some cases by as much as two orders of magnitude (Cr, Cu, and Zn), indicating that urba n fluvial source signals can ov erwhelm those from regional atmospheric sources. In estuarine systems, howev er, it appears that similar declines following the phase-out of leaded gasoline are not necessarily as rapid. Steding et al. (2000) used isotopic evidence to demonstrate the continued cycling of Pb in the San Francisco Bay estuary. In the southern arm of San Francisco Bay, which has an average depth of <2 m, Steding et al. (2000) found that isotopic compositions were essentially invariant, with 90% of the Pb derived from 1960s-1970s leaded gasoline. The authors attribut ed this to the limited hydraulic flushing and

remobilization of Pb from bottom sediments. In the northern arm of San Francisco Bay, although seasonal and decadal variations in Pb isotope composition were observed, mass balanceAX7-142 calculations indicate that only a small fraction of leaded gasoline fallout from the late 1980s has been washed out of the San Joaquin and S acramento rivers’ drainage basin by 1995 and consequently freshwater inputs remain a Pb sour ce to the bay (Steding et al., 2000). The authors suggest that the continuous sour ce of Pb from the river system s draining into the bay, coupled with benthic remobilization of Pb, indicates that historic gasoline deposi ts may remain in the combined riparian/estuarine system for decades.

AX7.2.2.3 Tracing the Fate and Transpor t of Lead in Aquatic Ecosystems The following section presents a generalized fram ework for the fate and transport of Pb in aquatic systems (Figure AX7-2.2.14). The prim ary source of Pb in natural systems is atmospheric deposition (Rickard and Nriagu, 197 8; U.S. Environmental Protection Agency, 1986a). Estimated median global atmospheric emission for anthropogenic and natural sources

are 332 Η 10 6 kg/year and 12 Η 10 6 kg/year, respectively (summa rized by Giusti et al., 1993). Inorganic and metallic Pb compounds are nonvolatile and will partition to airborne particulates or water vapors (Syracuse Res earch Corporation., 1999). Disper sion and deposition of Pb is dependent on the particle size (U.S. Envi ronmental Protection Agency, 1986a; Syracuse Research Corporation., 1999). More soluble form s of Pb will be removed from the atmosphere by washout in rain. In addition to atmospheric deposition, Pb may enter aquatic ecosystems through industrial or municipal wastewater effluents, storm water runoff, er osion, or direct point source inputs (e.g., Pb shot or accidental spills). Once in th e aquatic environment, Pb will partition between the various compartments of the system (e.g., dissolved phase, solid phase, biota). The movement of Pb between dissolved and particul ate forms is governed by factors such as pH, sorption, and biotransformation (see Section AX7.2.2.1). Lead bound to organic matter will settle to the bottom sediment layer, be assimilate d by aquatic organisms, or be resuspended in the water column. The uptake, accumulation, and toxici ty of Pb in aquatic organisms from water and sediments are influenced by various envi ronmental factors (e.g., pH, organic matter, temperature, hardness, bioavailability). Th ese factors are further described in Section AX7.2.3.4). The remainder of this section di scusses some methods for describing the distribution of atmospheric Pb in the aquatic environment.AX7-143 Figure AX7-2.2.14. Lead cycle in an aquatic ecosystem.

Sediment Core Dating and Source Tracing In addition to directly measuring Pb concentr ations in various aqua tic compartments (see Section AX7.2.3.3), it is useful to study the vertical distribution of Pb. Sediment profiling and core dating is a method used to determine the extent of accumulation of atmospheric Pb and provide information on potential anthropogenic sources. Sedime nt concentration profiles are typically coupled with Pb is otopic analysis. The isotope fingerprinting method utilizes measurements of the abundance of common Pb isotopes (i.e., 204 Pb, 206 Pb, 207 Pb, 208 Pb) to distinguish between natural Pb over geologic time and potential anthropogenic sources. Details

of this method were described in Sec tion AX7.1.2. The concentration of isotope 204 Pb has remained constant throughout time, while the ot her isotope species can be linked to various anthropogenic Pb sources. Typically, th e ratios or signatures of isotopes (e.g., 207 Pb: 206 Pb) are compared between environmental samples to indicate similarities or differences in the site being investigated and the potential known sources. Generally, Pb concentrations in sediment va ry with depth. For example, Chow et al. (1973) examined sediment Pb profiles in southe rn California. Lead concentrations were increased in the shallower sediment depths and co mparatively decreased at greater depths. These changes in sediment vertical concentration were attributed to higher anthropogenic Pb fluxesAX7-144 from municipal sewage, storm runoff, and at mospheric deposition. Similar experiments conducted throughout the United States have also s uggested an increase in Pb concentrations in the upper sediment layer concomitant with in creases in anthropogenic inputs (Bloom and Crecelius, 1987; Case et al., 1989; Rits on et al., 1999; Chillrud et al., 2003). Sediment Pb concentration profiles and isotope analysis have also been used to identify specific anthropogenic sources. For example, Flegal et al. (1987) used isotopic ratios to trace sources of Pb in mussels from Monterey Ba y, CA to a specific sl ag deposit. Several investigators have examined isot opic tracers to determine potential regional sources of Pb in eastern North America and the Great Lakes (Fle gal et al., 1989b; Graney et al., 1995; Blais, 1996). Water samples from Lake Erie and Lake Ontario were collected and analyzed. Lead isotope ratios ( 206 Pb: 207 Pb) from the lakes were compared to known ratios for Pb aerosols derived

from industrial sources in Canada and the United States and found to corr elate positively. This indicated that a majority of Pb in the lakes was derived from those industrial sources (Flegal et al., 1989b). Similarly, Gallon et al. (2006) used 206 Pb: 207 Pb ratios in sediment cores collected

from a 300 km transect in Canadian Shield headwa ter lakes to differentiate Pb contributions from

smelter emissions relative to Pb contributions from other anthropogenic inputs. The authors

were able to estimate the amounts of smelter-der ived Pb in sediment collected along the 300 km transect. Lead isotopes in sediment cores from Quebec and Ontario, Canada were also used to distinguish between the amount of Pb deposited from local Canadian sources (28.4 to 61.7%) and U.S. sources (38.3 to 71.6%) (Blais, 1996). Ex amination of Pb isotopes in sediment and suspended sediment in the St. Lawrence River we re used to identify pot ential anthropogenic Pb sources from Canada (Gobeil et al., 1995, 2005). Graney et al. (1995) used Pb isotope measurements to describe the differing historic sources of Pb in Lake Erie, Ontario and in Michigan. Temporal changes in Pb isotopic ratios were found to correspond to sources such as regional deforestation from 1860 through1890, coal combustion and or smelting through 1930, and the influence of leaded gaso line consumption from 1930 to 1980. The historic record of atmospheric Pb pollu tion has been studied to understand the natural background Pb concentration and the effects of Pb accumulation on ecosystems (Bindler et al., 1999; Renberg et al., 2000, 2002; Brännvall et al., 2001a,b). The most extensive work in this area has been conducted at pristine locations in Sw eden (Bindler et al., 1999). In this study, soil, sediment, and tree rings were sampled for Pb concentrations and isotopic analyses wereAX7-145 conducted on the soil samples. From this recor d, historic Pb concentrations and Pb accumulation

rates were estimated. Present da y concentrations in the forest soils ranged from 40 to 100 mg/kg, while a natural background concentr ation was estimated at <1 mg/kg. The authors were able to

model Pb concentrations for the past 6,000 years and also to proj ect Pb concentrations for the next 400 years, given an assumed atmos pheric deposition rate of 1 mg Pb m ! 2 /year. Models such as this are useful tools in determining the critical limits of metals in soils or sediments (Bindler et al., 1999; Renberg et al., 2002). Lead source association may also be asse ssed through retrospective measurements. Squire et al. (2002) used a time-se ries approach to evaluate the change in Pb in San Francisco Bay, CA from 1989 to 1999. This approach involved the use of detailed linear regression models and long-term monitoring data to determine change s in Pb concentrations and to identify events corresponding to those changes. Sediment and water samples were collected throughout the bay and combined with data on effluent discharges, urban runoff, atmospheric deposition, and river discharges. The authors identifi ed a 40% decline of Pb in the southern portion of the bay but found no change in the northern re ach. The decline was attributed to a reduction in wastewater source loadings over the previous decade.

AX7.2.2.4 Summary Lead is widely distributed in aquatic ecosystems, pre dominantly originating from atmospheric deposition or point source contribution. The fate and behavior of Pb in aquatic systems is regulated by physical and chemical f actors such as pH, salinity, sediment sorption, transformation, and uptake by aquatic biota. In th e United States, Pb concentrations in surface waters, sediments, and fish tissues range fr om 0.04 to 30 μg/L, 0.5 to 12,000 mg/kg, and 0.08 to 23 mg/kg, respectively. Atmospheri c sources are generally decreasi ng, as the United States has removed Pb from gasoline and other products. Howe ver, elevated Pb conc entrations remain at sites associated with mining wastes or wastewat er effluents. Since the 1986 Lead AQCD, much has been learned about the proces ses affecting Pb fate and tran sport. Detailed analyses are currently available (i.e., Pb isotope dating) to allow for constructi ng the history of Pb accumulation and identifying specific Pb contam inant sources. Continued source control along with examination of the physical and chemical properties will further a llow for the reduction of Pb concentrations throughout the United States.AX7-146 

### AX7.2.3 Aquatic Species Response/Mode of Action Recent advancements in understanding the resp onses of aquatic biota to Pb exposure are highlighted in this section. A summary of the co nclusions on the review of aquatic responses to Pb from the appropriate sections of the 1986 Lead AQCD, Volume II (U.S. Environmental Protection Agency, 1986a) and the subsequent co nclusions and recommen dations contained in the EPA staff review of that document (U.S. Environmental Protection Agency, 1990) are also provided. In addition, this sec tion summarizes research subs equent to the 1986 Lead AQCD on Pb uptake into aquatic biota, effects of Pb sp eciation on uptake, resistance mechanisms to Pb toxicity, physiological eff ects of Pb, factors that affect res ponses to Pb, and factors associated with global climate change. Areas of research that are not addr essed here include literature related to exposure to Pb shot or pellets and studies that exam ine human health-related endpoints (e.g., hypertension), which are described in other sections of this document.

AX7.2.3.1 Lead Uptake Lead is nutritionally nonessentia l and non-beneficial and is toxi c to living organisms in all of its forms (Eisler, 2000). Lead can bioaccumu late in the tissues of aquatic organisms through ingestion of food and water and adsorption fr om water (Vázquez et al., 1999; Vink, 2002) and subsequently lead to adverse effects if tissue levels are sufficiently high (see Section AX7.2.5). Recent research has suggested that due to the low so lubility of Pb in water, dietary Pb (i.e., lead adsorbed to sediment, particulate matter, and f ood) may contribute substan tially to exposure and toxicity in aquatic biota (B esser et al., 2005). Besser et al. (2004) exposed the amphipod Hyalella azteca to concentrations of Pb to evaluate th e influence of waterborne and dietary Pb exposure on acute and chronic toxi city. The authors found that acu te toxicity was unaffected by dietary exposure but that diet ary Pb exposure did contribute to chronic toxic effects (i.e., survival, growth, reproduction) in H. azteca . Field studies in ar eas affected by metal contamination (i.e., Clark Fork River, MO; Coeur d’Alene, ID) (Woodward et al., 1994, 1995; Farag et al., 1994) have also de monstrated the effects of diet ary metals on rainbow trout. However, there has been a debate on the impor tance of dietary exposure, as few controlled laboratory studies have been able to replicate the effects observed in the field studies (Hodson et al., 1978; Mount, 1994; Erikson, 2001) . This may be due to differe nces in the availability of Pb from the dietary sources used in laborat ory studies, differences in speciation, and/orAX7-147 nutritional characteristics of the Pb dosed diets. In many fiel d and laboratory studies, dietary exposure is rarily considered, but food provided to biota in these studie s adsorb metals from water. Therefore, both dietary and waterb orne exposure are occurring and both may be considered to play roles in elic iting the measured effects.

Lead concentrations in the ti ssues of aquatic organisms are generally higher in algae and benthic organisms and lower in higher trophic-l evel consumers (Eisler, 2000). Thus, trophic transfer of Pb through food chains is not expected (Eisler, 2000) . Metals are not metabolized; therefore, they are good integrat ive indicators of exposure in a quatic biota (Luoma and Rainbow, 2005). Metal uptake is complex, being influenced by geochemistry, route of exposure (diet and adsorption), depuration, and grow th (Luoma and Rainbow, 2005). This section discusses the factors affecting uptake of Pb by aq uatic biota and the state of current research in this area. As described in Section AX7.2.2.1, the solubili ty of Pb in water varies with pH, temperature, and ion concentra tion (water hardness) (Weber, 1993). Lead becomes soluble and bioavailable under conditions of low pH, or ganic carbon content, suspended sediment concentrations, and ionic concentrations (i.e., low Cd, Ca, Fe, Mn, Zn) (Eisler, 2000). Lead rapidly loses solubility above pH 6.5 (Rickard and Nriagu, 1978) and preci pitates out as Pb(OH) +

and PbHCO 3 + into bed sediments. However, at redu ced pH levels or ionic concentrations, sediment Pb can remobilize and potentially beco me bioavailable to aquatic organisms (Weber, 1993). The most bioavailable inorganic fo rm of Pb is divalent Pb (Pb 2+ ), which tends to be more readily assimilated by organisms than complexed forms (Erten-Unal et al., 1998). On the other hand, the low solubility of Pb salts restricts mo vement across cell membranes, resulting in less accumulation of Pb in fish in comparison to other metals (e.g., Hg, Cu) (Baatrup, 1991). The accumulation of Pb in aquatic organisms is , therefore, influenced by water pH, with lower pHs favoring bioavailability and accumulati on. For example, fish accumulated Pb at a greater rate in acidic lakes (pH = 4.9 to 5.4) than in more neutral lakes (pH = 5.8 to 6.8) (Stripp et al., 1990). Merlini and Pozzi (1977) found that pumpkinseed sunfish exposed to Pb at pH 6.0

accumulated three-times as much Pb as fish ke pt at pH 7.5. However, Albers and Camardese (1993a,b) examined the effects of pH on Pb uptake in aquatic plan ts and invertebrates in acidic (pH ~5.0) and nonacidic (pH ~6.5) constructed wetlands, ponds, and small lakes in Maine and Maryland. Their results suggested that low pH had little effect on the accumulation of metals byAX7-148 aquatic plants and insects and on the concentratio n of metals in the waters of these aquatic systems (Albers and Camardese, 1993a,b). Three geochemical factors that influence metal bioaccumulation in aquatic organisms include speciation, particulate metal form, and meta l form in the tissues of prey items (Luoma and Rainbow, 2005). Lead is typically pres ent in acidic aquatic environments as PbSO 4 , PbCl 4 , ionic Pb, cationic forms of Pb-hydr oxide, and ordinary hydroxide Pb(OH) 2 . In alkaline waters, common species of Pb include ani onic forms of Pb-carbonate (Pb(CO 3 )) and Pb(OH) 2 . Labile forms of Pb (e.g., Pb 2+ , PbOH + , PbCO 3 ) are a significant portion of the Pb inputs to aquatic systems from atmospheric washout. Particul ate-bound forms are more often linked to urban runoff and mining effluents (Eisle r, 2000). Little research has been done to link the complex concepts of chemical speciation and bioavail ability in natural systems (Vink, 2002). The relationship between the geochemistry of the unde rlying sediment and the impact of temporal changes (e.g., seasonal temperatures) to metal sp eciation are particularly not well studied (Vink, 2002; Hassler et al., 2004). Generally speaking, aquatic organisms exhi bit three Pb accumulation strategies: (1) accumulation of significant Pb concentrations with a low rate of loss, (2) excretion of Pb roughly in balance with availabil ity of metal in the environment, and (3) weak net accumulation due to very low metal uptake rate and no signi ficant excretion (Rainbow, 1996). Species that

accumulate nonessential metals such as Pb and th at have low rates of loss must partition it internally in such a way that it is sparingly available metabol ically. Otherwise, it may cause adverse toxicological effects (R ainbow, 1996). Aquatic organisms that exhibit this type of physiological response have been recommended fo r use both as environmental indicators of heavy metal pollution (Borgmann et al., 1993; Cast ro et al., 1996; Carter and Porter, 1997) and, in the case of macrophytes, as phytoremediators, because they accumulate heavy metals rapidly from surface water and sediment (Gavrilenko an d Zolotukhina, 1989; Simòes Gonçalves et al., 1991; Carter and Porter, 1997). Uptake experiments with aquatic plan ts and invertebrates (e.g., macrophytes, chironomids, crayfish) have shown steady in creases in Pb uptake with increasing Pb concentration in solution (Knowlton et al., 1983 ; Timmermans et al., 1992). In crayfish, the process of molting can cause a redu ction in body Pb concentrations, as Pb incorporated into the crayfish shell is eliminated (K nowlton et al., 1983). Vázquez et al. (1999) reported on the uptakeAX7-149 of Pb from solution to the extr acellular and intracellular compartm ents of 3 species of aquatic bryophytes. Relative to the 6 metals tested, Pb wa s found to accumulate to the largest degree in the extracellular compartments of all 3 bryophytes . The extracellular metals were defined as those that are incorporated into the cell wall or are found on the outer surface of the plasma membrane (i.e., adsorbed) (Vázquez et al., 1999). Intracellular me tals were defined as metals introduced into the cell through a meta bolically controlled process. Arai et al. (2002) examined the effect of growth on the uptake and elimination of trace

metals in the abalone Haliotos . They reported that older abal ones had generally lower whole body concentrations of heavy metals than did you nger, rapidly growing individuals. During the rapid growth of juveniles, the organism is less able to distinguish between essential (e.g., Zn), and nonessential metals (e.g., Pb). Once they re ach maturity, they develop the ability to differentiate these metals. Li et al. (2004) reported a similar response in zebra fish embryolarvae. Li et al. (2004) suggest ed that mature physiological sy stems are not developed in the embryo-larvae to handle elevated concentrations of metals. Therefore, metals are transported into the body by facilitated diffusion. Both the zebra fish and juvenile abalone demonstrate a rapid accumulation strategy followed by a low ra te of loss as described above. There are

insufficient data available to determine whet her this phenomenon is true for other aquatic organisms. Growth rates are generally thought to be an important consideration in the comparison of

Pb levels in individuals of the same species. The larger the indi vidual the more the metal content is diluted by body ti ssue (Rainbow, 1996). Once Pb is absorbed, it may sequester into varying parts of the organism. Calcium appears to have an important influence on Pb tran sfer. For example, Pb uptake and retention in

the skin and skeleton of coho salmon was reduced when dietary Ca was increased (Varanasi and Gmur, 1978). Organic Pb compounds tend to accumulate in lipids, and are taken up and accumulated in fish more readily than inorganic Pb compounds (Pattee and Pain, 2003). Given the complexities of metal uptake in natu ral systems, a model incorporating some of

the factors mentioned above is desirable. Th e EPA’s Environmental Re search Laboratory in Duluth, Minnesota developed a thermodynamic e quilibrium model, MINTEQ that predicts aqueous speciation, adsorption, precipitation, and/or dissolution of solids for a defined set of environmental conditions (MacDonald et al., 2002; Playle, 2004). Although not specificallyAX7-150 designed to model uptake, MINTEQ provides an indication of what forms of the metal are likely to be encountered by aquatic organisms by estima ting the formation of metal ions, complexation of metals, and the general bioa vailability of metals from environmental parameters. More recently, a mechanistic model centered on bi odynamics has been proposed by Luoma and Rainbow (2005) as a method of tying together geochemical influences, biological differences, and differences among metals to model metal bioaccumulation. The biodynamic model would be useful in determining the pot ential adverse effects on aquatic biota, which species are most useful as indicators of metal effects, and how ecosystems may change when contaminated by metals. Two prominent models examine trace me tal bioavailability and its link to effects (Hassler et al., 2004). These include the free i on activity model (FIAM) and the biotic ligand model (BLM). Specific information on these mode ls, including their lim itations, is provided in

Section AX7.2.1.3. These models are useful fo r advancing our understanding of how metal uptake occurs in aquatic organisms and how uptake and toxic effects are linked.

Bioconcentration Factors (BCF) BCFs for Pb are reported for va rious aquatic plants in Ta ble AX7-2.3.1. The green alga Cladophora glomerata is reported as having the highest BCF (Keeney et al., 1976). Duckweed ( Lemna minor ) exhibited high BCF values ranging from 840 to 3560 depending on the method of measurement (Rahmani and Sternberg, 1999). Duckw eed that was either previously exposed or not exposed to Pb was exposed to a single dos e of Pb-nitrate at 5000 μg/L for 21 days. Duckweed that was previously exposed to Pb rem oved 70 to 80% of the Pb from the water, while the previously unexposed duckweed removed 85 to 90%. Both plant groups were effective at removing Pb from the water at sublethal levels. BCFs for Pb are reported for various i nvertebrates in Table AX7-2.3.2. BCFs for freshwater snails were 738 for a 28-day e xposure (Spehar et al., 1978) and 1,700 for a 120-day exposure (Borgmann et al., 1978). Other reported values for invertebrates included a BCF of 1930 for the scud during a 4-day exposure (MacL ean et al., 1996), and BCF s of 499 and 1120 for the caddis fly and stonefly, respectively, in 28-d ay exposures (Spehar et al., 1978). In a 28-day exposure, midge larvae were reported with a BCF of 3670 (Timme rmans et al., 1992).AX7-151 Table AX7-2.3.1. Bioconcentration Factors for Aquatic Plants

BCF Species Test Conditions Reference

840 to 2700 (measured digestion) Duckweed ( Lemna minor ) 21 days, Pb-nitrate Rahmani and Sternberg (1999)

1150 to 3560 (measured solution) Duckweed 21 days, Pb-nitrate Rahmani and Sternberg (1999)

16,000 to 20,000 Alga (Cladophora glomerata ) not specified Keeney et al. (1976)

Table AX7-2.3.2. Bioconcentration Fact ors for Aquatic Invertebrates

BCF Species Test Conditions Reference 738 Snail ( Physa integra ) 28 days, Pb-nitrate Spehar et al. (1978)

1700 Snail ( Lymnaea palustris ) 120 days, Pb-nitrate Borgmann et al. (1978)

499 Caddis fly ( Brachycentrus sp. ) 28 days, Pb-nitrate Spehar et al. (1978)

1120 Stonefly ( Pteronarcys dorsata ) 28 days, Pb-nitrate Spehar et al. (1978)

1930 Scud ( Hyalella azteca ) 4 days, Pb-chloride MacLean et al. (1996)

3670 Midge larvae ( Chironomus riparius ) 28 days Timmermans et al. (1992)

BCFs for freshwater fish were 42 and 45 for brook trout and bluegill, respectively (Holcombe et al., 1976; Atchis on et al., 1977). Although no BCFs have been reported for amphibians, Pb-nitrate was reported to accumulate ma inly in the ventral skin and in the kidneys of frogs (Vogiatzis and Loumbourdis, 1999).

Bioconcentration factors and bioaccumulation fa ctors (BAFs) are not necessarily the best predictors of tissue concentration levels given exposure concentration levels (Kapustka et al., 2004). The role of homeostatic mechanisms is a major consideration in tissue concentrations found in exposed biota. Similarly, measuring BC Fs and BAFs in organisms may not accurately reflect how metals are treated within the orga nisms (e.g., partitioning to specific organelles,AX7-152 sequestering to organ tissues). Therefore, they are not recommended for use in conducting metal risk assessments (Kapustka et al., 2004).

AX7.2.3.2 Resistance Mechanisms Detoxification Mechanisms Detoxification includes the biological proces ses by which the toxic qualities, or the probability and/or severity of harmful effects, of a poison or toxin are reduced by the organism. In the case of heavy metals, this process fre quently involves the sequestration of the metal, rendering it metabolically inactive. Recent resear ch into heavy metal detoxification in aquatic biota has focused on several phys iological and biochemical mechanisms for detoxifying Pb.

This section examines these mechanisms and the abil ity of plants, protists, invertebrates, and fish to mitigate Pb toxicity.

Plants and Protists Deng et al. (2004) studied the uptake and tran slocation of Pb in wetland plant species surviving in contaminated sites. They found that all plants tended to sequester significantly larger amounts of Pb in their roots than in their shoots. Deng et al. (2004) calculated a translocation factor (TF), the am ount of Pb found in the shoots divided by the amount of Pb found in the root system, and found that TFs ranged from 0.02 to 0.80. Concentrations of Pb in

shoots were maintained at low levels and vari ed within a narrow ra nge. Deng et al. (2004) observed that plants grown in Pb-contaminated sites usually contained higher concentrations than the 27 mg/kg toxicity threshold established for plants by Beckett a nd Davis (1977). Some of the wetland plants examined by Deng et al. (2004) also accumu lated high concentrations of metals in shoot tissues; however, these metals were assumed to be detoxified (metabolically unavailable), as no toxic response to these elevated concentrations was observed. Deng et al. (2004) suggested that this ability is likely related to discrete internal metal detoxification

tolerance mechanisms. Phytochelatins are thiol-cont aining intracellular metal-bi nding polypeptides that are produced by plants and protists in response to excessive uptake of heavy metals (Zenk, 1996). Phytochelatins are synthesized by the enzyme ph ytochelatin synthase th at is activated by the presence of metal ions and uses glutathione as a substrate. When phytoc helatins are synthesizedAX7-153 in sufficient amounts to chelate the metal ion, th e enzyme is deactivated (Morelli and Scarano, 2001). Morelli and Scarano (2001) studied phytochela tin synthesis and stability in the marine

diatom Phaeodactylum tricornutum in the presence of Pb. They found that when metal exposure was alleviated, significant cellular Pb-phytochela tin complex content was lost. Their findings support a hypothesis of vacuolar ization proposed for higher plan ts (Zenk, 1996), in which metalphytochelatin complexes are activel y transported from the cytosol to the vacuole, where they undergo rapid turnover. Zenk (1996) suggested that the complex dissociates, and the metal-free peptide is subsequently degraded. Morelli a nd Scarano (2001) proposed concomitant occurrence of phytochelatin synthesis and release during me tal exposure, as a coincident detoxification

mechanism in P. tricornutum

.

Aquatic Invertebrates Like plants and protists, aquatic animal s detoxify Pb by preventing it from being metabolically available, though their mechanis ms for doing so vary. Invertebrates use lysosomal-vacuolar systems to sequester and pro cess Pb within glandular cells (Giamberini and Pihan, 1996). They also accumulate Pb as de posits on and within skeletal tissue (Knowlton et al., 1983; Anderson et al., 1997; Boisson et al., 2002), and some can efficiently excrete Pb (Vogt and Quinitio, 1994; Prasuna et al., 1996). Boisson et al. (2002) used radiot racers to evaluate the transfer of Pb into the food pathway

of the starfish Asterias rubens as well as its distributi on and retention in various body compartments. Boisson et al. (2002) monitored Pb elimination after a single feeding of Pbcontaminated molluscs and found that Pb was sequ estered and retained in the skeleton of the starfish, preventing it from being metabolically available in other tissues. Elimination (as percent retention in the skeleton) was found to follow an exponential time course. Elimination was rapid at first, but slowed after 1 week, and eventually stabilized, implying an infinite biological half-life for firmly bound Pb. Results of radiotracer tracking sugge st that Pb migrates within the body wall from the organic matrix to th e calcified skeleton. From there, the metal is either absorbed directly or adsorbed on newl y produced ossicles (small calcareous skeletal structures), where it is efficiently retained as mineral deposition and is not metabolically active (Boisson et al., 2002).AX7-154 AbdAllah and Moustafa (2002) studied the Pb st orage capability of organs in the marine

snail Nerita saxtilis . Enlarged electron-dense vesicles and many granules were observed in digestive cells of these snails a nd are suggested to be the site of storage of detoxified metals.

N. saxtilis were found to be capable of concentrat ing Pb up to 50 times that of surrounding marine water without exhibiting signs of histopathologic change s. This ability has been

attributed to chelation with various bioc hemical compounds, such as thionine (forming metallothionine) (Rainbow, 1996), or complexa tion with carbonate, forming lipofuchsin (AbdAllah and Moustafa, 2002). Granules observed in lysosomal residual bodies were presumed to be the result of Pb accumulation. The presen ce of large vacuoles and residual bodies were indicative of the fragmentation phase of diges tion, suggesting that Pb was also processed chemically in the digestive cells. The podocyte cells of the pericardial gland of biva lves are involved in the ultrafiltration of the hemolymph (Giamberini and Pihan, 1996). A microanalytical study of the podocytes in Dreissena polymorpha exposed to Pb revealed lysosomal-v acuolar storage/processing similar to that in the digestive cells of Nerita saxtilis. The lysosome is thought to be the target organelle for trace metal accumulation in various organs of bivalves (Giamberini and Pihan, 1996). Epithelial secretion is the principal det oxification mechanism of the tiger prawn Penaeus

monodon . Vogt and Quinitio (1994) found that Pb granules tended to accumulate in the epithelial cells of the an tennal gland (the organ of excretio n) of juveniles exposed for 5 and 10 days to waterborne Pb. The metal is depos ited in vacuoles belonging to the lysosomal system. Continued deposition leads to the fo rmation of electron-dens e granules. Mature granules are released from the cells by apocrine secretion in to the lumen of the gland, and presumably excreted through the nephridopore (i .e., the opening of the antennal gland). Apocrine secretion is predominant, so that as granules form, they are kept at low levels. Excretion was also found to be a primary and e fficient detoxification mechanism in the shrimp Chrissia halyi (Prasuna et al., 1996). Crayfish exposed to Pb have been shown to concentrate the metal in their exoskeleton and exuvia through adsorption processes. More than 80% of Pb found in exposed crayfish has been found in exoskeletons (Knowlton et al., 1983; A nderson et al., 1997). Following exposure,

clearance is most dramatic from the exoskeleton. The result of a 3-week Pb-clearance study with red swamp crayfish Procambarus clarkia , following a 7-week exposure to 150 μg Pb/L, showedAX7-155 an 87% clearance from the exoskeleton due, in part , to molting. Other organs or tissues that take up significant amounts of Pb include the gills , hepatopancreas, muscle, and hemolymph, in decreasing order. These parts cleared >50% of accumulated Pb over the 3-week clearance period, with the exception of the hepatopancreas. The hepato pancreas is the organ of metal storage and detoxification, although the molecular mechanisms of me tal balance in crayfish have yet to be extensively inves tigated (Anderson et al., 1997).

Fish

Most fish use mucus as a first line of de fense against heavy metals (Coello and Khan, 1996). In fish, some epithelia are covered with extracellular mucus secreted from specialized cells. Mucus contains glyc oproteins, and composition varies among species. Mucosal glycoproteins chelate Pb, and se ttle, removing the metal from the water column. Fish may secrete large amounts of mucus when they come into contact with pot ential chemical and biochemical threats. Coello and Khan (1996) in vestigated the role of externally added fish mucus and scales in accumulating Pb from water, and the relationship of th ese with the toxicity of Pb in fingerlings of green sunfish, goldfish and largemouth bass. The authors compared trials in which fish scales from black sea bass ( Centropristis striata ) and flounder ( Pseudopleuronectes

americanus ) and mucus from largemouth bass were added to green sunfish, goldfish, and largemouth bass test systems and to reference test systems. On exposure to Pb, fish immediately started secreting mucus from epidermal cells in va rious parts of the body. Metallic Pb stimulated filamentous secretion, mostly from the ventrolatera l areas of the gills, while Pb-nitrate stimulated

diffuse molecular mucus secretion from all over the body. The addition of largemouth bass mucus significantly increased the LT 50 (the time to kill 50%) for green sunfish and goldfish exposed to 250 mg/L of Pb-nitrate. In contrast, Tao et al . (2000) found that mucus reduced the overall bioavailability of Pb to fish but that the reduction was insignificant. Coello and Khan (1996) found that scales were more significant in reducing LT 50 than mucous. Fish scales can accumulate high concentrations of metals, includi ng Pb, through chelation with keratin. Scales were shown to buffer the pH of Pb-nitrate in solu tion and remove Pb from water after which they

settled out of the water column. Addition of scales to test water made a ll species (green sunfish, goldfish, and largemouth bass) more tolerant of Pb.AX7-156 Summary of Detoxifiction Processes Mechanisms of detoxification vary among aqua tic biota and include processes such as translocation, excretion, chelation, adsorption, vacuolar storage, and deposition. Lead

detoxification has not been studied extensively in aquatic organism s, but existing results indicate the following:

• Protists and plants produce intr acellular polypeptides that fo rm complexes with Pb (Zenk, 1996; Morelli and Scarano, 2001).

• Macrophytes and wetland plants that thrive in Pb-contaminated regions have developed translocation strategies for tolerance and detoxification (K nowlton et al., 1983; Deng et al., 2004).

• Some starfish (asteroids) se quester the metal via mineral deposition into the exoskeleton (Boisson et al., 2002).

• Species of mollusc employ lysosomal-vacuolar sy stems that store and chemically process Pb in the cells of their digestive and pericardi al glands (Giamberini and Pihan, 1996; AbdAllah and Moustafa, 2002).

• Decapods can efficiently excrete Pb (Vogt a nd Quinitio, 1994; Giamberini and Pihan, 1996) and sequester metal through adsorption to the exoskeleton (Knowlton et al., 1983).

• Fish scales and mucous chelate Pb in the water column, and potent ially reduce visceral exposure.

Avoidance Response Avoidance is the evasion of a perceived th reat. Recent research into heavy metal avoidance in aquatic organisms has looked at dose-re sponse relationships as well as the effects of

coincident environmental factors. Preference/avoidance response to Pb has not been extensively studied in aquatic organisms. In particular , data for aquatic invertebrates is lacking. Using recent literature, this section exam ines preference-avoidance responses of

invertebrates and fish to Pb and some other environmental gradients.

Aquatic Invertebrates Only one study was identified on avoidance re sponse in aquatic invertebrates. Lefcort et al. (2004) studied the avoidance behavior of the aquatic pulmonate snail Physella columbiana from a pond that had been polluted with heavy me tals for over 120 years. In a Y-maze test, first generation P. columbiana from the contaminated site avoided Pb at 9283 μg/L (p < 0.05) andAX7-157 moved toward Pb at 6255 μg/L (p < 0.05). It is thought that attraction to Pb at certain elevated

concentrations is related to Pb neuron-stimulating properties (Lef cort et al., 2004). These results

are consistent with those from si milar studies. Control snails fr om reference sites, and first and second-generation snails from contaminated site s were capable of detecting and avoiding heavy metals, although the first genera tion was better than the second generation, and the second was better than the controls at doing so. This sugge sts that detection and avoidance of Pb is both genetic and environmentally based for P. columbiana. Lefcort et al. (2004) observed heightened sensitivity to, and avoidance of, heavy metals by the snails when metals where present in

combination.

Aquatic Vertebrates Steele et al. (1989) studie d the preference-avoidan ce response of bullfrog ( Rana

catesbeiana ) to plumes of Pb-contaminated water following 144-h exposure to 0 to 1000 μg Pb/L. In this laboratory experi ment, tadpoles were exposed to an influx of 1000 μg Pb/L at five different infusion rates (i.e., volumes per unit tim e into the test system). Experiments were videotaped and location data from the tank were used to assess respons e. No significant differences were seen in pref erence-avoidance responses to Pb in either nonexposed or previously exposed animals. In a similar subsequent st udy, Steele et al. (1991) studied preference-avoidance response to Pb in American toad ( Bufo americanus ) using the same exposure range (0 to 1000 μg Pb/L). B. americanus did not significantly avoid Pb, and behavioral stress responses were not observed. The results do not indicate whether the tadpoles were capable of perceiving the contaminant. Lack of avoidance may indicate insufficient perception or the lack of physiolo gical stress (Steele et al., 1991). The olfactory system in fish is involved in their forming avoidance response to heavy metals (Brown et al., 1982; Svecevi č ius, 1991). It is generally t hought that behavi oral avoidance of contaminants may be a cause of reduced fi sh populations in some water bodies, because of disturbances in migration and distribution patterns (Svecevi č ius, 2001). Unfortunately, avoidance of Pb by fish has not been studied as extensively as for other heavy metals (Woodward et al., 1995). Woodward et al. (1995) studi ed metal mixture avoidan ce response in brown trout ( Salmo trutta ), as well as the added effect s of acidification. A 1-fold (1 Η ) mixture containedAX7-158 1.1 μg/L Cd, 12 μg/L Cu, 55 μg/L Zn, and 3.2 μg/L Pb (all metals were in the form of chlorides). Avoidance was quantified as time spen t in test water, trip time to te st water, and number of trips.

Brown trout avoided the 1 Η mixture as well as the 0.5 Η , 2 Η , 4 Η , and 10 Η mixtures, but not the

0.1 Η mixture. Reduced avoidance was observed at higher concentrations (4 Η and 10 Η ). The authors proposed that the reduced avoidance re sponse was due to impaired perception due to injury. These responses are typical of other fish species to individual metals of similar concentrations (Woodward et al., 1995). This st udy does not conclusively indicate which of the metals in the mixture may be causing the avoi dance response. Howeve r, given the neurotoxic effects of Pb, impaired perception is a likely response of Pb-exposed fish. When test water was reduced in pH from 8 to 7, 6 to 5, brown trout avoidance increased, but with no significant difference between metal treatments and controls. However, in the 1 Η metal mixture treatment, brown trout made fewer tr ips into the test water chamber at the lower pHs (Woodward et al., 1995). This response may be related to an increased abundance of Pb cations at lower pH values in the test system. Scherer and McNicol (1998) investigated th e preference-avoidance responses of lake whitefish ( Coregonus clupeaformis ) to overlapping gradients of li ght and Pb. Whitefish were found to prefer shade in untreated water. Lead concentrations under illumination ranged from 0 to 1000 μg/L, and from 0 to 54,000 μg/L in the shade. Under uniform illumination, Pb was avoided at concentrations above 10 μg/L, but avoidance behavior l acked a dose-dependent increase over concentrations ranging from 10 to 1000 μg Pb/L. Avoidance in shaded areas was strongly suppressed, and whitefish only avoided Pb at concentra tions at or above 32,000 μg/L.

Summary of Avoidance Response In summary, of those aquatic organisms studi ed, some are quite adept at avoiding Pb in aquatic systems, while others seem incapable of detecting its presence. Snails have been shown to be sensitive to Pb and to avoid it at high c oncentrations. Conversely, anuran (frog and toad) species lack an avoidance response to the metal. Fish avoidance of many chemical toxicants has been well established, and it is a dominant subl ethal response in po lluted waters (Svecevi č ius, 2001). However, no studies have been located sp ecifically examining avoidance behavior for Pb in fish. Environmental gradients, such as light and pH, can alter preference-avoidance responses.AX7-159 AX7.2.3.3 Physiological Effects of Lead This section presents a review of the physiological effects and functional growth responses associated with the exposure of aquati c biota to Pb. Physiological effects of Pb on aquatic biota can occur at the biochemical, cellula r, and tissue levels of organization and include inhibition of heme formation, adverse effects to blood chemistry, and decreases in enzyme levels. Functional growth responses resulting from Pb exposure include changes in growth patterns, gill binding affiniti es, and absorption rates.

Biochemical Effects Lead was observed to have a gender-selec tive effect on brain endocannabinoid (eCB) (e.g., 2-arachidonylglyc erol [2-AG] and N -arachidonylethanolamine [AEA]) levels in fathead

minnow Pimephales promelas (Rademacher et al., 2005). Cannabi noids, such as eCB, influence locomotor activity in organisms. Increased levels of cannabinoids have been shown to stimulate locomotor activity and decreased levels slow locomotor activity (Sa ñudo-Peña et al., 2000). Male and female fathead minnows were exposed to 0 and 1000 μg/L of Pb. Female minnows in the control group contained signifi cantly higher levels of AEA and 2-AG compared to males. At a concentration of 1000 μg Pb/L, this patte rn reversed, with males showing significantly higher levels of AEA in the brain than fema les (Rademacher et al., 2005). After 14-days exposure to the 1000 μg Pb/L trea tment, significantly higher levels of 2-AG were found in male fathead minnows, but no effect on 2-AG levels in females was observed (Rademacher et al., 2005). Lead acetate slightly inhibited 7-ethoxyresorufin- o -deethylase (7-EROD) activity in Gammarus pulex exposed for up to 96 h to a single toxicant concentration (EC 50 ) (Kutlu and Susuz, 2004). The exact concentration used in the study was not reported. The EROD enzyme is required to catalyze the c onjugation and detoxification of toxic molecules and has been proposed as a biomarker for contaminant exposure. The authors believe more detailed studies are required to confirm EROD as a biomarker for Pb exposure. The enzyme group alanine transferases (ALT) has been suggested as a bi oindicator/biomarker of Pb stress (Blasco and Puppo, 1999). A negative correlation was obs erved between Pb accumulation and ALT concentrations in the gills and soft body of Ruditapes philippinarum exposed to 350 to 700 μg/L of Pb for 7 days (Blasco and Puppo, 1999).AX7-160 Studies have identified ALAD in fish and amphi bians as a useful indicator of Pb exposure (Gill et al., 1991; Nakagawa et al., 1995a,b). ALAD catalyzes the formation of hemoglobin and early steps in the synthesis of protoporphyrin (Gill et al., 1991; Nakagawa et al., 1995b). The absence of an inhibitory effect on this enzyme following exposure to cadmium, copper, zinc, and mercury suggests that this enzyme reacts specif ically to Pb (Johansson-Sjöbeck and Larsson, 1979; Gill et al., 1991). A 0% decrease in ALAD activity was reported in common carp ( Cyprinus carpio ) exposed to a Pb concentration of 10 μg/L for 20 days (Nakagawa et al., 1995b). The recovery of ALAD activity after exposur e to Pb has also been examined in carp (Nakagawa et al., 1995a). After 2-week exposur e to 200 μg Pb/L, ALAD activity decreased to approximately 25% of value reported for controls (Nakagawa et al., 1995a). Fish removed from

the test concentration after 2 w eeks and placed in a Pb-free envi ronment recovered slightly, but ALAD activity was only 50% of the controls even after 4 weeks (Nakagawa et al., 1995a). Vogiatzis and Loumbourdis (1999) exposed the frog ( Rana ridibunda ) to a Pb concentration of 14,000 μg/L over 30 days and a 90% decrease in ALAD activity was observed in the frogs.

Blood Chemistry

Numerous studies have examined the effects of Pb exposure on blood chemistry in aquatic biota. These studies have primarily used fish in acute and chronic exposures to Pb concentrations ranging from 100 to 10,000 μg/L. Decreased erythroc yte, hemoglobin, and hemocrit levels were observed in rosy barb ( Barbus puntius ) during an 8-week exposure to 126 μg/L of Pb-nitrate (Gill et al., 1991).

No difference was found in red blood cell c ounts and blood hemoglobin in yellow eels ( Anguilla anguilla ) exposed to 0 and 300 μg/L of Pb for 30 days (Santos and Hall, 1990). The number of white blood cells, in the form of ly mphocytes, increased in the exposed eels. The

authors concluded this demonstrates the lasting action of Pb as a toxicant on the immune system (Santos and Hall, 1990). Significant decreases in red blood cell counts and volume was reported in blue tilapia ( Oreochromis aureus ) exposed to Pb-chloride at a concentration of 10,000 μg/L for 1 week (Allen, 1993). Blood components, such as plasma glucose, total plasma protei n, and total plasma cholesterol, were unaffected in yellow eels exposed to 300 μg/L of Pb for 30 days (Santos and Hall, 1990). Effects on plasma chemistry were observed in Oreochromis mossambicus exposedAX7-161 to 0, 18,000, 24,000, and 33,000 μg/L of Pb (Ruparelia et al., 1989). Significant decreases in plasma glucose (hypoglycemic levels) were reported at concentrations of 24,000 and 33,000 Φ g Pb/L after 14 and 21 days of exposure, and at 18,000 μg Pb/L after 21 days of exposure (Ruparelia et al., 1989). Plasma cholesterol levels dropped significantly in comparison to controls after 14 days of exposur e to 33,000 μg Pb/L and in all test concentrations after 21 days of exposure (Ruparelia et al., 1989). Similarly, concentrations of blood serum protein, albumin, and globulin were identified as bioind icators of Pb stress in carp ( Cyprinus carpio ) exposed to

Pb-nitrates at concentrations of 800 a nd 8000 μg Pb/L (Gopal et al., 1997).

Tissues In fish, the gills serve as an active site for ion uptake. Recent studies have examined the competition between cations for binding sites at the fish gill (e.g., Ca 2+ , Mg 2+ , Na + , H + , Pb 2+ ) (MacDonald et al., 2002; Rogers and W ood, 2003, 2004). Studie s suggest that Pb 2+ is an antagonist of Ca 2+ uptake (Rogers and Wood, 2003, 2004). MacDonald et al. (2002) proposed a gill-Pb binding model that assumes Pb 2+ has a ∃ 100 times greater affinity for binding sites at the fish gill than other cations. More toxicity studies are required to quantify critical Pb burdens that

could be used as indicators of Pb toxicity (Niyogi and Wood, 2003).

Growth Responses A negative linear relationship was observ ed in the marine gastropod abalone ( Haliotis ) between shell length and muscle Pb concentrations (Arai et al., 2002). Abalones were collected from two sites along the Japanese coast. Haliotis discus hannai were collected from along the coast at Onagawa; Haliotis discus were collected from along the co ast at Amatsu Kominato. The authors did not report significant differences between the two sampling sites. From samples collected at Onagawa, Pb concentrations of 0.03 and 0.01 μg/g were associated with abalone shell lengths of 7.7 cm (3 years old) and 12.3 cm (6 years old), respectively. From samples collected at Amatsu Kominato, Pb concentratio ns of 0.09 and 0.01 μg/g were associated with abalone shell lengths of 3.9 cm (0 years old) a nd 15.3 cm (8 years old), re spectively (Arai et al., 2002). The authors theorized that young ab alones, experiencing rapid growth, do not discriminate between the uptake of essential a nd nonessential metals. However, as abalones grow larger and their rate of growth decreases, they increasin gly favor the uptake of essentialAX7-162 metals over nonessential metals. This is demonstr ated by the relatively cons istent concentrations of Cu, Mn, and Zn that were reported fo r the abalone samples (Arai et al., 2002).

Other Physiological Effects

Increased levels of Pb in water were found to increase fish production of mucus: excess mucus coagulates were observed over the entire body of fishes. Buildup was particularly high around the gills, and in the worst cases, interfer ed with respiration and resulted in death by anoxia (Aronson, 1971; National Resear ch Council of Canada., 1973).

AX7.2.3.4 Factors That Modify Organism Response to Lead A great deal of research has b een undertaken recently to be tter understand the factors that modify aquatic organism response to metals incl uding lead. A discussion of research on the many factors that can modify aquatic organism re sponse to Pb is provided in this section.

Influence of Organism Age and Size on Lead Uptake and Response It is generally accepted that Pb accumulation in living organisms is controlled, in part, by metabolic rates (Farkas et al., 2003 ). Metabolic rates are, in-turn, controll ed by the physiological conditions of an organism, including such factor s as size, age, point in reproductive cycle, nutrition, and overall health. Of these physiolo gical conditions, size and age are the most commonly investigated in relation to heavy metal uptake. This section reviews recent research focusing on relationships between body size, age, and Pb accumulation in aquatic invertebrates

and fish.

Invertebrates MacLean et al. (1996) investigated bioaccumula tion kinetics and toxicity of Pb in the amphipod Hyalella azteca . Their results indicated that body size did not greatly influence Pb

accumulation in H. azteca exposed to 50 or 100 μg/L of PbCl 2 for 4 days. Canli and Furness (1993) found similar results in the Norway lobster Nephrops norvegicus exposed to 100 μg/L of Pb(NO 3 ) 2 for 30 days. No significa nt sex- or size-related differences were found in concentrations of Pb in the tissue. The highest tissue burden was found in the carapaces (42%).

Several studies have determined that Pb can bind to the exoskeleton of invertebrates andAX7-163 sometimes dominate the total Pb accumulated (K nowlton et al., 1983). This adsorption of Pb to

the outer surface of invertebrates can result in strong negative relationships for whole-body Pb

concentration as a function of body mass (i.e., concentrations decr ease rapidly with increased body size and then stabilize) (MacLean et al., 1996). Drava et al. (2004) investigated Pb concen trations in the muscle of red shrimp Aristeus

antennatus from the northwest Mediterranean. L ead concentrations ranged from 0.04 to 0.31 μg/g dw. No significant relationships between size and Pb concentration in A. antennatus were found, and concentrations were not related to reproductive status. Arai et al. (2002) analyzed abalones ( Haliotis ) at various life stages from coastal regions of Japan. They investigated growth effects on the uptake and elimination of Pb. Results indicated a significant negative linear relati onship between age, shell length and Pb

concentrations in muscle tissue. The relationshi p was consistent despite habitat variations in Pb concentrations between the study sites, sugge sting that Pb concentr ations changed with growth in the muscle tissue of test specimens and implying that abalone can mitigate Pb exposure as they age.

Fish Douben (1989) investigated the effects of body size and age on Pb body burden in the stone loach ( Noemacheilus barbatulus L.). Fish were caught dur ing two consecutive springs from three Derbyshire rivers. Results indicated that Pb burden increased slightly with age. Similarly, Köck et al. (1996) found that concentrati ons of Pb in the liver and kidneys of Arctic char ( Salvelinus alpinus ) taken from oligotrophic alpine lakes were posi tively correlated with age. It has been suggested that fish are not able to eliminate Pb completely, and that this leads to a stepwise accumulation from year to year (Köck et al., 1996). In contra st, Farkas et al. (2003) found a negative relationship between Pb concentr ations and muscle and gill Pb concentrations

in the freshwater fish Abramis brama . Fish were taken from a low-contaminated site and contained between 0.44 and 3.24 μg/g Pb dw. Negative correlations between metal

concentration and fish size in low-contaminated wa ters likely results from variations in feeding rates associated with developmental stages. This hypothesis is consistent with the fact that in low-contaminated waters, feeding is the main r oute of uptake and feedi ng rates decrease with development in fish (Farkas et al., 2003).AX7-164 In summary, relationships between age, size, and Pb body burden in aquatic invertebrates and fish are interspecifically variable and depend on many envi ronment-related variables (e.g., exposure) (Farkas et al., 2003).

Genetics There are few studies documenting the e ffects of Pb on organismal and population genetics, although rapid advances in biotechnology have prompted recent research in this area (Beaty et al., 1998). There are two principal eff ects that sublethal exposure to a contaminant can have on the genetics of an organism and/or population: (1) a contaminant may influence selection by selecting for certain phenotypes that enable populati ons to better cope with the chemical; or (2) a contaminant can be genotoxic, meaning it can produce alterations in nucleic acids at sublethal exposure concen trations, resulting in changes in hereditary characteristics or DNA inactivation (Shugart, 1995). Laboratory studies have shown that exposure to Pb 2+ at 10 mg/mL in blood produces chromosomal aberrations (i.e., deviations in the normal structure or number of chromosomes) in some organisms (Ces tari et al., 2004). Eff ects of genotoxicity and toxin-induced selection do not preclude one another, and may act together on exposed populations. This section reviews Pb genotoxicity and the effects of Pb-induced selection in aquatic populations.

Selection Evidence for genetic selection in the natura l environment has been observed in some aquatic populations exposed to metals (Rand et al., 1995; Beaty et al., 1998; Duan et al., 2000; Kim et al., 2003). Because tolerant individua ls have a selective advantage over vulnerable individuals in polluted environmen ts, the frequency of tolerance genes will increase in exposed populations over time (Beaty et al ., 1998). Several studies have shown that heavy metals can alter population gene pools in aquatic invertebrate s. These changes have resulted in decreased genetic diversity and are thought to be a potentia l source of population instability (Duan et al., 2000; Kim et al., 2003). Kim et al. (2003) investigated genetic di fferences and population structuring in the gastropod Littorina brevicula from heavy-metal polluted and unpolluted environments. Organisms from polluted sites contained a m ean of 1.76 μg Pb/g, while organisms fromAX7-165 unpolluted sites contained 0.33 μg Pb/g. They found significant differences in haplotypes between the test groups and allelic di versity was significantly lower among L. brevicula from polluted regions. In contrast, Ya p et al. (2004) performed a sim ilar experiment with the greenlipped mussel Perna viridis ; they found that mussels from contaminated sites containing between 4 and 10 μg Pb/g, as well as other heavy meta ls, exhibited a higher percentage of polymorphic loci and excess heterozygosity compared to those from uncontaminated sites. The higher level of genetic diversity was attribut ed to greater environmental hete rogeneity (i.e., variation due to pollution gradients) in contaminat ed sites (Yap et al., 2004). Duan et al. (2000) i nvestigated amphipod ( Hyalella azteca ) selective mortality and genetic structure following acute exposure to Pb (5.47 mg/L Pb(NO 2 ) 2 ) as well as exposure to other heavy metals. They found that genetic differentiation consiste ntly increased among survivors from the original population, supporti ng the hypothesis that heavy metals, including Pb, have the potential to alter th e gene pools of aquatic organisms.

Genotoxicity Lead exposure in water (50 μg/L) over 4 weeks resulted in DNA strand breakage in the

freshwater mussel Anodonta grandis (Black et al., 1996), although hi gher concentrations (up to 5000 μg/L) did not result in significant breakag e by the end of the study period. These results suggest that a threshold effect for DNA damage and repair exists, where DNA repair only occurs once a certain body exposure level has been reach ed. More recently, Cestari et al. (2004) observed similar results in neotropical fish ( Hoplias malabaricus ) that were fed Pb-contaminated food over 18, 41, and 64 days. Lead body burdens in H. malabaricus were approximately 21 μg

Pb 2+ /g. Results indicated that exposure to Pb significantly increased the frequency of chromosomal aberrations and DNA damage in ki dney cell cultures, although when assessed at the end of the longer exposure period s, aberrations were less common.

Environmental Biological Factors Environmental factors that are biological in origin can alter the av ailability, uptake and toxicity of Pb to aquatic organisms. These factors can be grouped into living and non-living constituents. For example, living organisms ma y sequester Pb from the water column, reducing the availability and toxicity of the metal in the water column to other biota, thus reducingAX7-166 potential toxic effects in othe r organisms. Non-living organi c material (e.g., components of sloughed-off scales, mucus, carcasses, and othe r decomposing, humic material) can similarly combine with Pb from the water column, rendering it unavailable. This section will review the literature on biological environmen tal factors and their influence on the bioavailability, uptake, and toxicity of Pb. Van Hattum et al. (1996) studied the influence of abiotic va riables, including DOC on Pb concentrations in freshwater isopods ( Proasellus meridianus and Asellus aquaticus ). They found that BCFs were significantly negatively correlat ed with DOC concentrations. Thus, as DOC concentrations increased, BCFs decreased in P. meridianus and A. aquaticus , indicating that DOC acts to inhibit the availability of Pb to these isopods. Kruatrachue et al. (2002) invest igated the combined effects of Pb and humic acid on total chlorophyll content, growth rate, multiplicati on rate, and Pb uptake of common duckweed.

When humic acid was added to the Pb-nitrat e test solutions ( 50, 100, and 200 mg Pb(NO 3 ) 2 / L), toxicity of Pb to duckweed was decreased. The addition of humic acid to the Pb-nitrate solution increased the pH. The authors suggested that there was a proton dissociation from the carboxyl group in the humic acid that complexed with Pb, resu lting in a decrease in free Pb ions available to the plant. Schwartz et al. (2004) collected natural organic matter ( NOM) from several aquatic sites across Canada and investigated the effects of NOM on Pb toxicity in rainbow trout ( Oncorhynchus mykiss ). They also looked at toxicity e ffects as they related to the optical properties of the various NOM samples. The re sults showed that NOM in test water almost always increased LT 50 and that optically dark NOM tended to decrease Pb toxicity more than did optically light NOM in rainbow trout. In summary, non-living constituents of biologi cal origin in the environment have been shown to reduce Pb availability a nd, therefore, toxicity in some aquatic organisms. It is generally thought that this occurs through complexa tion or chelation processe s that take place in

the water column.AX7-167 Physicochemical Environmental Factors

This section reviews the literature on physicochemical environm ental factors and their influence on the bioavailability, uptake, and toxic ity of Pb in aquatic organisms. These factors are discussed with regard to their infl uence individually and in combination. Studies generally agree that as pH increases , the toxicity of Pb decreases (Horne and Dunson, 1995b; MacDonald et al., 2002) . As pH decreases, Pb becomes more soluble and more readily bioavailable to aqua tic organisms (Weber, 1993). Significantly lower survival, decreased hatching success, slower development, and in creased egg mass and la rval mortality were observed in Jefferson salamanders ( Ambystoma jeffersonianum ) and wood frogs ( Rana sylvatica ) exposed to Pb at a pH of 4.5 versus a pH of 5.5 (Horne and Dunson, 1995b). Contradictory results have been reported for invertebrates. Over a 96-h exposure peri od, mortality increased with decreasing pH for the bivalve Pisidium casertanum , while pH-independent mortality was reported for gastropods and crustacea under si milar exposure conditions (Mackie, 1989). Cladocerans ( C. dubia ) and amphipods ( H. azteca ) were also more sensitive to Pb toxicity at pH 6 to 6.5 than at higher pH levels (Schubauer-Ber igan et al., 1993). Lead was 100 times more toxic to the amphipod Hyalella azteca at a pH range of 5.0 to 6.0 (Mackie, 1989) than at a pH range of 7.0 to 8.5 (Schubauer-Berigan et al., 19 93). Lead was also more toxic to fathead minnows at lower pH levels (Sc hubauer-Berigan et al., 1993). The influence of pH on Pb accumulation has also been observed in sediments. Accumulation of Pb by the isopod Asellus communis was enhanced at low pH

, after a 20-day exposure to Pb-contaminated sediment s (Lewis and McIntosh, 1986). In A. aquaticus , temperature increases were found to be more im portant than increased pH in influencing Pb accumulation (Van Hattum et al., 1996). Increased water temperature was also found to reduce Pb uptake fluxes in green microalga ( Chlorella kesslerii ) (Hassler et al., 2 004). Lead and zinc body concentrations in Asellus sp. were found to vary marked ly with seasonal temperature changes, with greater concentrations present in spring and summer (Van Hattum et al., 1996). Acute and chronic toxicity of Pb increases w ith decreasing water hardness, as Pb becomes more soluble and bioavailable to aquatic orga nisms (Horne and Dunson, 1995a; Borgmann et al., 2005). There is some evidence that water hardness and pH work together to increase or decrease the toxicity of Pb. Jefferson salamanders expos ed to Pb for 28 days at low pH and low water hardness experienced 50% mortality, while exposu re to Pb at high pH and high water hardnessAX7-168 resulted in 91.7% survival (H orne and Dunson, 1995a). Exposur e to Pb at high pH and low water hardness or low pH and high water ha rdness resulted in 75 and 41.7% survival, respectively (Horne and Dunson, 1995a). Si milar results were reported for Jefferson salamanders during a 7-day exposure and wood fr ogs during 7- and 28-day exposures (Horne and Dunson, 1995c). In some cases, water hardne ss and pH in the absence of Pb have been shown to affect survival adversely. Mean acute survival of wood frogs and Jefferson salamanders exposed to low pH and low water hardness, in the absence of Pb, was 83.3 and 91.7%, respectively. Mean chronic survival of wood frogs and Jefferson salamanders exposed to low pH and low water hardness, in the absence of Pb, was 79.2 and 41.7%, respectively (Horne and Dunson, 1995c). High Ca 2+ concentrations have been shown to pr otect against the toxic effects of Pb (Sayer et al., 1989; MacDonald et al., 2002; Hassler et al., 2004; Rogers and Wood, 2004). Calcium affects the permeability and integrity of cell membranes and intracellular contents (Sayer et al., 1989). As Ca 2+ concentrations decr ease, the passive flux of ions (e.g., Pb) and

water increases. At the lowest waterborne Ca 2+ concentration (150 μmol/L), Pb accumulation in juvenile rainbow trout ( Oncorhynchus mykiss ) branchials significantly increased as Pb

concentration in water increased (Rogers and Wood, 2004). At higher Ca 2+ concentrations, Pb accumulation did not significantly increase with Pb concentration in water. This result demonstrates the protective effects of waterborne Ca 2+ and supports the suggestion that the Ca 2+ component of water hardness determines the toxi city of Pb to fish (Rogers and Wood, 2004). Rogers and Wood (2004) reporte d that the uptake of Ca 2+ and Pb 2+ involves competitive inhibition of apical entry at lanthanum-sensitive Ca 2+ channels and interfer ence with the function

of the ATP-driven baso-lateral Ca 2+ pump. High mortality was reported in brown trout ( Salmo

trutta ) fry exposed to Pb at a waterborne Ca 2+ concentration of 20 μmol/L, while negligible mortality was reported at the same Pb concentration but at a waterborne Ca 2+ concentration of 200 μmol/L (Sayer et al., 1989). Adverse effect s to mineral uptake a nd skeletal development

were observed in the latter test group (Sayer et al., 1989). The bioavailability of Pb and other metals that can be simultaneously extracted in sediments may be modified through the role of acid volatile sulf ide (AVS) under anoxic conditions (Tessier and Campbell, 1987; Di Toro et al., 1992; Casas and Crecelius, 1994). The term AVS (iron sulfide is an example) refers to the fraction of the sediment that consists of aAX7-169 reactive pool of solid-phase sulfide. This phase is available to bind divalent metals that then become unavailable for uptake by aquatic biota. The models proposed by Di Toro et al. (1992) and Casas and Crecelius (1994) predict that when the molar ratio of simultaneously extractable metals (SEM) to AVS in sediments is less than one, the metals will not be bioavailable due to complexation with available sulfide. Salinity is an important modi fying factor to metal toxici ty. Verslycke et al. (2003) exposed the estuarine mysid Neomysis integer to individual metals , including Pb, and metal mixtures under changing salinity. At a salinity of 5%, the reported LC 50 for Pb was 1140 μg/L (95% CL = 840, 1440 μg/L). At an increased salinity of 25 l , the toxicity of Pb was substantially reduced (LC 50 = 4274 μg/L [95% CL = 3540, 5710 μg/L]) (Verslycke et al., 2003). The reduction in toxicity was attributed to increased complexation of Pb 2+ with Cl ! ions.

Nutritional Factors The relationship between nutrition and Pb toxic ity has not been thoroughly investigated in aquatic organisms. In fact, algae species are the only aquatic organisms to have been studied fairly frequently. Although nutrients have been found to have an impact on Pb toxicity, the mechanisms involved are poorly understood. It is unclear whether the relationship between nutrients and toxicity comprises organismal nut rition (the process by wh ich a living organism assimilates food and uses it for growth and for replacement of tissues), or whether nutrients have interacted directly with Pb, inhibiting its metabol ic interaction in the organism. This section

reviews the little information that has been gathered from studies documenting apparent Pbnutrition associations in aquatic organisms. Jampani (1988) looked at the impact of various nutrients (i.e., sodium acetate, citric acid, sodium carbonate, nitrogen, and phosphates) on reduc ing growth inhibition in blue-green algae ( Synechococcus aeruginosus ) exposed to 200 mg Pb/L. E xposure to this Pb treatment concentration caused 100% mortality in algae. Results indicated that additional nitrogen, phosphates, and some carbon sources, including sodium acetate, citric acid and sodium carbonate, all protected the algae fr om Pb toxicity. Algae that had been starved prior to the experiment were found to be significantly more sensitive to Pb exposure. Glucose was the only

nutrient tested that did not have a si gnificant impact on Pb toxicity in S. aeruginosus . In a similar study by Rao and Reddy (1985) on Scenedesmus incrassatulus , nitrogen, phosphate andAX7-170 carbon sources (including glucose), all had protectiv e effects, and reduced Pb toxicity at 300 and 400 mg Pb/L. Both studies proposed similar hypot heses regarding nutrient-Pb mechanisms that led to reduced toxicity. One hypothe sis was that the nutrients were ab le to reverse toxic effects. The second hypothesis was that the nutrients in teracted directly with Pb, in some way sequestering the metal so as to inhibit its metabolic interaction with the organism (Rao and Reddy, 1985; Jampani, 1988). Rai and Raizada (1989) investig ated the effects of Pb on nitrate and ammonium uptake as well as carbon dioxide and nitrogen fixation in Nostoc muscorum over a 96-h period. Test specimens were exposed to 10, 20, and 30 mg Pb/L. At 20 mg Pb/L, nitrate uptake was inhibited by 64% after 24 h and by 39% after 96 h. Am monium uptake was inhibited, and similarly,
inhibition decreased from 72% i nhibition after 24 h to 26% inhibi tion after 96 h of exposure. Carbon dioxide fixation and nitroge nase activity followed similar pa tterns, and results indicated that Pb exposure can affect th e uptake of some nutrients in N. muscorum

. Adam and Abdel-Basset (1990) studied the e ffect of Pb on meta bolic processes of Scenedesmus obliquus . They found that nitrog enase activity was inhibi ted by Pb nitrate, but enhanced by Pb-acetate. As photosynthetic pr oducts and respiratory substrates, carbohydrate and lipid levels were altered by Pb. Above 30 mg/L of Pb-nitrate, both macronutrients were reduced. However, Pb-acetate was found to incr ease carbohydrate levels. Results suggest that

Pb can have an effect on macronutrients in S. obliquus and that effects may vary depending on the chemical species. Simòes Gonçalves et al. (1991) studied the impact of light, nutrients, air flux, and Pb, in various combinations, on growth inhibition in the green algae Selenastrum capricornutum

. Results indicated that at lo wer Pb concentrations (<0.207 mg/L) and increased nutrient concentrations, algae release more exudates that form inert comp lexes with Pb anions in the water. This suggests that S. capricornutum can use exudates as a protection and that this protective mechanism depends on nutrient supply. These results are consis tent with those of Capelo et al. (1993), who investigated uptak e of nitrogen and phosphorus in the algae Selenastrum capricornutum over time in the absence and pr esence of 0.207 mg Pb/L. They found that the presence of Pb ha d no significant influence on the assimilation of nitrogen and phosphorus. However, they did find that in the presence of Pb, algae released higher concentrations of organics with Pb-chelating groups.AX7-171 Amiard et al. (1994) investigated the impact on soft tissue Pb concentrations of various feeding regimes on oysters ( Crassostrea gigas ) during their spat rearing. They fed test groups of C. gigas different amounts of Skeletonema costatum and additional natural phytoplankton grown

in test solutions. Results showed that size and food intake both negatively correlated with metal concentrations in soft tissue. The authors hypothesi zed that this relationshi p was due in part to a diluting effect of the food. In summary, nutrients affect Pb toxicity in t hose aquatic organisms that have been studied. Some nutrients seem to be capable of reducing toxicity, though the mechanisms have not been well established. Exposure to Pb has not been shown to reduce nutrien t uptake ability, though it has been demonstrated that Pb exposure may l ead to increased production and loss of organic material (e.g., mucus and other complex organic ligands) (Cap elo et al., 1993).

Interactions with Other Pollutants

Most of the scientific literature reviewed in this section considered how Pb and other elements combine to affect uptak e and exert toxicity. Research on the interactions of Pb with complexing ligands and other physic al and biological factors was more t horoughly discussed in Section AX7.2.3.4. Predicting the response of organi sms to mixtures of chemicals is difficult (Norwood et al., 2003). There are two schools of thought on addressing chemical mixtures in toxicology. The first focuses on the combined mode of action of the individual mixture substances while the other focuses on the organism response to the mixture as being additive, or some deviation from additive (synergistic or an tagonistic). The mode of action model assumes

that each of the individual substances has sim ilar pharmacokinetics as well as mode of action. Approaches to modeling combined mode of ac tion include: 1) joint independent action for

toxicants with different modes of action; and 2) joint similar mode of action. Within the scientific literature the deviations from additivity approach can be confusing as some authors report concentration additivity, while others report effect additivit y. The concentration additivity

model assumes that the sum of the concentrations will result in a level of effect similar to the simple sum of the effects observed if each chemi cal were applied separately (e.g., Herkovits and Perez-Coll, 1991). Toxic units (TU) may be used to describe the concentration additivity results for mixtures (e.g., Hagopian-Schlekat et al., 2001). Effects additivity sugges ts that the level of

effect of a mixture will be the sum of the effects of each chemical used separately. Thus, theAX7-172 separate concentration-effects models for each chem ical are used to predict individual effects and the sum of these predictions is then compared to the actual effect of the mixture. Deviations

from the mixture effect are then classified as less than addi tive (antagonistic) or more than additive (synergistic) (see Piegorsch et al ., 1988; Finney 1947 for further information). The complexity of metal mixture interact ions as different metal concentrations, environmental conditions (e.g., temperature, pH), and other factors can cause marked changes in the effects observed (Norwood et al., 2003). In describing Pb interactions with other elements, the different approaches to modeling mixture toxi city are considered. Specific reference to known Pb-metal interactions and implications on Pb uptake and toxicity will also be made in

each of the studies below. Less than additive or antagonistic interactions can reduce metal bioavailability when metals are present in combination, and may lead to reduced potential for toxicity (Hassler et al., 2004). A number of elements act in an antagonistic fashion with P b. For example, Pb is a wellknown antagonist to Ca 2+ (Niyogi and Wood, 2004; Hassler et al., 2004), which is an essential element, required for a number of physiological pr ocesses in most organisms. Lead ions have an

atomic structure similar to Ca 2+ and can be transported either actively or passively across cell membranes in place of Ca 2+ . An example of this interaction was reported by Behra (1993a,b)

where Pb was shown to activate calm odulin reactions in rainbow trout ( O. mykiss ) and sea mussel ( Mytilus sp.) tissues in the absence of cal cium. Calmodulin (CaM) is a major intracellular calcium receptor and regulates th e activities of numerous enzymes and cellular processes. Allen (1994) reported that Pb can replace calcium in body structures (e.g., bones, shells); replace zinc in ALAD, which is requ ired for heme biosynthesis; and react with sulfhydryl groups, causing conformation protein distortion and scission of nucleic acids (Herkovits and Perez-Coll, 1991). Lead is also a known antagonist to Mg 2+ , Na + , and Cl ! regulation in fish (Ahern and Morris, 1998; Rogers and Wood, 2003, 2004; Niyogi and Wood, 2004). Li et al. (2004) repor ted on the interaction of Pb 2+ with Cd 2+ in the context of adsorption from solution by Phanerochaete chrysosporium , a filamentous fungus. The authors found that cadmium uptake decreased with increasi ng concentration of Pb ions with Pb 2+ outcompeting

Cd 2+ for binding sites. Hassler et al. (2004) reported that in the presence of copper (Cu 2+ ), there was a significantly higher rate of interna lization of Pb in the green algae Chlorella kesserii . It wasAX7-173 suggested that Cu 2+ may have affected organism phys iology through the disruption of cell membrane integrity. This would allow increased cation (i.e., Pb 2+ ) permeability and, therefore, substantially increase internalization of Pb and result in effects that were more than additive (synergistic). Hagopian-Schlekat et al. (2001) ex amined the impact of individual metals and complex metal mixtures containing Cd, Cu, Ni, Zn, and Pb to the estuarine copepod Amphiascus

tenuiremis . The copepods were exposed to metal-spik ed sediment and pore water. The mixed metal sediment toxicity tests demonstrat ed greater than additive toxicity to A. tenuiremis . It was postulated that the synergism observed was due to two or more metals affecting the same biological function. Herkovits and Perez-Coll (1991) exposed Bufo arenarum larvae to various

Pb and zinc concentrations in solution. At low zinc concen trations, (2:1 Pb:Zn ratio), a synergistic toxic effect was obs erved in the frog larvae relativ e to the effects observed from exposure to the individual metals and at higher zi nc concentrations. Enhanced Pb toxicity was

attributed to the interference of Pb with cellular activities due to binding with sulfhydryl polypeptides and nucleic acid phos phates (Herkovits and Pere z-Coll, 1991). Allen (1994) reported on the accumulation of numerous metals a nd ions into specific tissues of the tilapia

Oreochromis aureus . Tilapia exposed to low concentra tions of Pb and mercury (both at 0.05 mg/L) had significantly higher concentrations of Pb in internal organs than those fish exposed to Pb alone. Similarly, low concentrati ons of cadmium with low concentrations of Pb caused increased uptake of Pb in certain or gans (e.g., liver, brain, and caudal muscle). Lead has been shown to complex with Cl ! in aquatic systems. For example, Verslycke et al. (2003) exposed the estuarine mysid Neomysis integer to six different me tals, including Pb, and a combined metal mixture under changing sali nity conditions. At a salinity of 5%, the reported LC 50 for Pb was 1140 μg/L (840, 1440 μg/L). At an increased salinity of 25 l , the toxicity of Pb was s ubstantially reduced (LC 50 = 4274 μg/L [3540, 5710 Φ g/L]) (Verslycke et al., 2003). This reduction in toxicity was a ttributed to the increas ed concentration of Cl ! ion due to increased salinity, in that it complexed with divalent Pb in the te st system. Exposure of N. integer to Pb in combination with the other five metals (Hg, Cd, Cu, Zn, Ni) resulted in roughly strictly additive toxicity (Verslycke et al., 2003). Long et al. (2006) performed a critical review of the uses of mean sediment quality guideline quotients (mSQGQs) in assessing the toxi c effects of contaminant mixtures (metals and organics) in sediments. This approach has b een used in numerous su rveys and studies sinceAX7-174 1994. mSQGQs are useful to risk assessors but their inherent limitations and underlying assumptions must be fully unde rstood (see Section AX7.2.1.5).

Summary of Interactions With Other Pollutants Norwood et al. (2003) reported th at in a review and reinterp retation of published data on the interactions of metals in binary mixtures (n = 15 studies), antagonistic (6) and additive interactions (6) were the most common for Pb. The complexity of the interactions and possible modifying factors makes determining the impact of even binary metal mixtures to aquatic biota difficult (Norwood et al., 2003; Playle, 2004). The two most commonly reported Pb-element

interactions are between Pb and calcium and between Pb and zinc . Both calcium and zinc are essential elements in organisms and the interactio n of Pb with these ions can lead to adverse effects both by increased Pb uptake and by a decrease in Ca and Zn required for normal

metabolic functions.

AX7.2.3.5 Factors Associated with Global Climate Change It is highly unlikely that Pb has any in fluence on generation of ground-level ozone, depletion of stratospheric ozone, global warming, or other indicator s of global climate change. Lead compounds have relatively short residence times in the atmosphere, making it unlikely that they will reach the stratosphere, and they do not absorb infrared radiation, making them unlikely to contribute to stratospheric ozone depletion or global warming. Also, these compounds are unlikely to have a significant interaction with gro und-level nitrogen oxides or volatile organic compounds, thus precluding genera tion of ground-level ozone. Approached from another viewpoint, climat e change can have a major impact on the fate/behavior of Pb in the environment and, th erefore, can subsequently alter organism or ecosystem responses. For example, changes in temperature regime (Q10 rule), changes in precipitation quantity and quality (e.g., acidic deposition) may influence fate, transport, uptake, and bioavailability of Pb (Syracuse Research Corporation., 1999).

AX7.2.3.6 Summary

There have been a number of advancements in the understanding of Pb behavior in the environment and its impact on aquatic organisms since 1986. In particular , greater knowledge ofAX7-175 factors that influence Pb accumulation in aquati c organisms, mechanisms of detoxification and avoidance of Pb, and greater underst anding of the interactions of Pb in aquatic systems. All of these factors require some consideration when a ttempting to determine the potential for exposure and subsequent response of aquatic species to lead. The following section provides a review on the current understanding of lead and its impacts on aquatic biota.



### AX7.2.4 Exposure/Response of Aquatic Species This section outlines and highlights the critic al recent advancements in the understanding of the toxicity of Pb to aquatic biota. The sect ion begins with a review of the major findings and conclusions from the 1986 Lead AQCD (U.S. En vironmental Protection Agency, 1986a). The following sections summarize the research conducted since 1986 on determining the

concentrations of Pb that cause th e effects discussed in Section 7.2.3. Lead exposure may adversely affect organisms at different levels of organization, i.e., individual organisms, populations, communities, or ecosystems. Generally, however, there is insufficient information available for single mate rials in controlled studies to permit evaluation of specific impacts on higher levels of organizatio n (beyond the individual organism). Potential effects at the population level or higher are, of necessity, extrapolated from individual level studies. Available population, community, or ecos ystem level studies are typically conducted at sites that have been contaminated or adversely affected by multiple stressors (several chemicals alone or combined with physical or biological stressors). Theref ore, the best documented links

between lead and effects on the environment are w ith effects on individual organisms. Impacts on aquatic ecosystems are discussed in Section 7.2.5 and Annex AX7.2.5.

AX7.2.4.1 Summary of Conclusions From the Previous Criteria Document The 1986 Lead AQCD (U.S. Environmental Prot ection Agency, 1986a) reviewed data in

the context of the sublethal eff ects of lead exposure. The do cument focused on describing the types and ranges of lead exposures in ecosystems likely to adversely impact domestic animals. As such, the criteria document did not provide a co mprehensive analysis of the effects of lead to most aquatic primary producers, consumers, and decomposers. For the aquatic environment, general reviews of the effects of lead to alga e, aquatic vertebrates, and invertebrates were undertaken. A summary of these reviews is provided below.AX7-176 Algae The 1986 Lead AQCD (U.S. Environmental Protection Agency, 1986a) reported that some algal species (e.g., Scenedesmus sp.) were found to exhibit physiological changes when exposed to high lead or organolead concentratio ns in situ. The observed changes included increased numbers of vacuoles, deformations in cell organelles, and increased autolytic activity.

Increased vacuolization was assumed to be a tolerance mechanism by which lead was

immobilized within cell vacuoles.

Aquatic Vertebrates The 1986 Lead AQCD (U.S. Environmental Protection Agency, 1986a) reported that hematological and neurological responses were the most commonly reported effects in aquatic

vertebrates. These effects include red blood cell destruction and inhi bition of the enzyme ALAD, required for hemoglobin synthesis. At hi gh lead concentrations, neurological responses included neuromuscular distortion, anorexia, muscle tremor, and spinal curvature (e.g., lordosis). The lowest reported exposure concentration causing either hematological or neurological effects was 8 μg/L (U.S. Environmental Protection Agency, 1986a).

Aquatic Invertebrates

Numerous studies were cited on the effects of lead to aqua tic invertebrates in the 1986 Lead AQCD (U.S. Environmental Protection Agency, 1986a). In general, lead concentrations in aquatic invertebrates were found to be correlated cl osely with concentrations in water rather than

food. Freshwater snails were found to accumulate l ead in soft tissue, ofte n in granular bodies of precipitated lead. Mortality and reproductive effects were reported to begin at 19 μg Pb/L for the

freshwater snail Lymnea palutris and 27 μg Pb/L for Daphnia sp. The review of the NAAQS for Pb (U.S. E nvironmental Protection Agency, 1990) made only one recommendation reported in the sections of the 1986 Lead AQCD dealing with effects to aquatic biota. This was the need to consider the impact of water hardness on Pb bioavailability and toxicity, to be consistent with the recommendations of the AWQC for the protection of aquatic life (U.S. Envi ronmental Protection Agency, 1985).AX7-177 AX7.2.4.2 Recent Studies on Effects of Lead on Primary Producers Using literature published si nce the 1986 Lead AQCD (U.S . Environmental Protection Agency, 1986a), this section examines the toxicity of Pb (individually and in metal mixtures) to algal and aquatic plant growth, its effects on metabolic processe s (e.g., nutrient uptake), and its impact on primary productivity in natural systems.

Toxicity of Lead to Algae The toxicity of Pb to algal growth has been investigated for a number of species including Chlorella vulgaris , Closterium acerosum , Pediastrum simplex , Scenedesmus quadricauda , Scenedesmu obliquus , Syneschoccus aeruginosus , and Nostoc muscorum (Jampani, 1988; Rai and Raizada, 1989; Adam and Abdel-Basset, 19 90; Fargašová, 1993; Bilgrami and Kumar, 1997). Study durations ranged from 7 to 20 days and Pb-nitrate was the most commonly used

form of Pb. Effect s to algal growth ( Chlorella vulgaris , Closterium acerosum , Pediastrum simplex , Scenedesmus quadricauda ), ranging from minimal to complete inhibition, have been reported at Pb concentrations between 100 and 200,000 μg/L (Jampani, 1988; Bilgrami and Kumar, 1997). Most studies report the percent inhibition in test groups compared to controls

rather than calculati ng the LOEC, NOEC, or EC 50 values. Clinical signs of Pb toxicity include the deformation and disintegration of algae cel ls and a shortened exponential growth phase (Jampani, 1988; Fargašová, 1993). Other effects of Pb block the pathways that lead to pigment synthesis, thus affecting photosynthesis, the cel l cycle and division, and ultimately result in cell death (Jampani, 1988). From the studies reviewed, Closterium acerosum is the most sensitive alga species tested (Bilgrami and Kumar, 1997). Exposure of these algae to 1000 and 10,000 μg/L as lead nitrate for 6 days resulted in cell gr owth that was 52.6 and 17.4%, respec tively, of controls (Bilgrami and Kumar, 1997). Chlorella vulgaris , Pediastrum simplex , and Scenedesmus quadricauda were also exposed to Pb-nitrate in this study. Compared to controls, cell growth at 1000 and 10,000 μg Pb-nitrate/L was 65.3 and 48.7%, 64.5 and 42.7%, and 77.6 and 63.2%, respectively (Bilgrami and Kumar, 1997). Scenedesmus quadricauda exhibited a similar magnitude of effects when exposed to lead (Pb 2+ ) for 20 days at 0, 5500, 11,000, 16,500, 22,000, 27,500, and 33,000 μg/L (Fargašová, 1993). This study reported an EC 50 for growth inhibition at 13,180 μg/L (95% CI: 10,190, 14,620). Decreased cell number, but incr eased cell size, wasAX7-178 observed in Selenastrum capricornutum 7 exposed to lead (Pb 2+ ) at 207.2 μg/L and a Q/V (flux of air [Q] divided by volume of the culture [V]) of 4.7 Η 10 ! 3 sec ! 1 for 9 days (Simòes Gonçalves et al., 1991). The Q/V is a measure of culture growth where an increase in the Q/V ratio indicates growth. The pigment concentration per cell decreased with exposure to Pb, so while the algae cells were larger, they were less hea lthy (Simòes Gonçalves et al., 1991). Growth rates were not reported, making comparison with other studies difficult. High Pb concentrations were required to elicit effects in Nostoc muscorum and Scenedesmus aeruginosus (Jampani, 1988; Rai and Raizada, 1989). Following 15-day exposures, test groups exposed to 10,000, 20,000, a nd 30,000 μg Pb/L experienced growth rates that were 90.5, 76.9, and 66.7% of the controls (Rai and Raizada, 1989). Synechococcus aeruginosus experienced little inhibiti on of growth from exposure to Pb-nitrate up to a concentration of 82,000 μg/L (Jampani, 1988). At a test concentration of 100,000 μg/L, complete inhibition of growth was observed, and at a concentration of 200,000 μg/L, algae failed to establish a single colony (Jampani, 1988). Scenedesmus obliquus are quite tolerant to the

effects of Pb-nitrate and Pb-acetat e on growth. Algae exposed to Pb-nitrate or Pb-acetate up to 180,000 μg/L had higher cell numbers than controls (Adam and Abdel-Basset, 1990). Exposure to the highest concentration of 300,000 μg/L Pb-n itrate or Pb-acetate resulted in cell numbers that were 81 and 90% of the controls, re spectively (Adam and Abdel-Basset, 1990). Lead in combination with other metals (e.g., Pb and Cd, Pb and Ni, etc.) is generally less toxic than exposure to Pb al one (Rai and Raizada, 1989). Nostoc muscorum exposed to

chromium and Pb in combination demonstrated bette r growth than when exposed to either of the metals alone (Rai and Raizada, 1989). Antagonis tic interaction was obser ved in the exposure of

Nostoc muscorum to Pb and nickel in combination (R ai and Raizada, 1989). When applied separately, these metals demonstr ated different levels of toxic ity; however, in combination, they exerted similar effects (Rai and Raizada, 1989). More information on toxic interactions of Pb with other metals is provided in Section AX7.2.3.5.

7 The species name Selenastrum capricornutum has been changed to Pseudokirchneriella subcapitata . The former species name is used in this report.AX7-179 Aquatic Plants The toxicity of Pb to aquatic pl ant growth has been studied using Spirodela polyrhiza , Azolla pinnata , and Lemna gibba (Gaur et al., 1994; Gupta a nd Chandra, 1994; Miranda and Ilangovan, 1996). Test durations ranged from 4 to 25 days and test c oncentrations ranged between 49.7 and 500,000 μg Pb/L (Gaur et al., 1994; Miranda and Ilangovan, 1996). Research on aquatic plants has focused on the effects of Pb on aquatic plant grow th and chlorophyll and protein content. Of the species reviewed here , the effects of Pb on aqua tic plant growth are most pronounced in Azolla pinnata (Gaur et al., 1994). An EC 50 of 1100 μg/L was reported for A. pinnata exposed to Pb-nitrate for 4 days. S. polyrhiza exposed to Pb-nitrate under the same test conditions had a reported EC 50 for growth of 3730 μg/L (Gaur et al., 1994). Lemna gibba

was shown to be the least sens itive plant species to Pb: si gnificant growth inhibition was reported at concentrations of 200,000 μg/L or greate r after 25 days of expos ure to concentrations of 30,000, 50,000, 100,000, 200,000, 300,000, or 500,000 μg/L (M iranda and Ilangovan, 1996). The maximum growth rate for L. gibba was observed at 10 days of exposure. After this point, the growth rate declined in controls and te st concentrations (Mir anda and Ilangovan, 1996). Clinical signs of Pb toxicity include yellowi ng and disintegration of fronds, reduced frond size, and chlorosis (Gaur et al., 1994; Miranda and Il angovan, 1996). Toxicity results suggest that effects to growth from Pb exposure occur in a dose-dependent manner (Gaur et al., 1994).

Effects of Lead on Metabolic Processes Algal and aquatic plant metabolic processes ar e variously affected by exposure to Pb, both singularly and in combination with other metals. Lead adversely affects the metabolic processes of nitrate uptake, nitrogen fixa tion, ammonium uptake, and carbon fixation at concentrations of 20,000 μg Pb/L or greater (Rai a nd Raizada, 1989). Lead in co mbination with nickel has an antagonistic effect on nitrogen fi xation and ammonium uptake, but a synergistic effect on nitrate uptake and carbon fixation (Rai and Raizada, 1989). Lead in comb ination with chromium has an antagonistic effect on nitrate up take, but it has a synergistic effect on nitrogen fixation, ammonium uptake, and carbon fixati on (Rai and Raizada, 1989). Lead effects on nitrate uptake in Nostoc muscorum (μg NO 3 /μg Chl a ) were greatest after 24 h, when exposure to 20,000 μg/L reduced nitr ate uptake by 64.3% compared to controls.AX7-180 Nitrate uptake reported after 48, 72, a nd 96 h was reduced by 30.0, 37.5, and 38.9%, respectively, compared to controls (Rai a nd Raizada, 1989). Lead in combination with chromium, both at a test c oncentration of 20,000 μg/L, demons trated antagonistic effects on nitrate uptake. Compared to controls, nitr ate uptake was reduced by 52.4, 30, 25, and 22.2% at 24, 48, 72 and 96 h, respectively (Rai and Rai zada, 1989). The greatest effect on uptake occurred at 24 h when, compared to controls , a 52.4% reduction was re ported in the test

concentration. Lead and nickel in combinati on at test concentrations of 20,000 and 1000 μg/L, respectively, resulted in a grea ter reduction of nitrate uptake than Pb alone at 48, 72, and 96 h (Rai and Raizada, 1989). After 24, 48, and 72 h of Pb exposure at 20,000 μg/L, nitrogenase activity (nmol C 2 H 4 / μg protein/hr) in Nostoc muscorum was reduced by 39.3, 61.8, and 14.1%, respectively, compared to controls (Rai and Raizada, 1989). A concen tration of 207.2 μg Pb/L had little effect on nitrogen or phosphorus assimilation in Selenastrum capricornutum over 7 days (Capelo et al., 1993). An antagonistic effect on nitrogenase activity was generally reported for Nostoc

muscorum exposed to Pb in combination with nickel at 20,000 and 1,000 μg/L, respectively (Rai and Raizada, 1989). Compared to cont rols, nitrogenase activ ity was reduced by 42.9, 32.7, and 13.6% at 24, 48, and 72 h, respectively (Rai a nd Raizada, 1989). Lead and chromium, both administered at a concentration of 20,000 μg/L, ha d a synergistic impact on nitrogenase activity

in Nostoc muscorum . Nitrogenase activity in the te st group was reduced by 60.7, 60, and 50% compared to the controls at 24, 48, and 72 h, respectively (Rai and Raizada, 1989). Lead-induced inhibition of ammonium uptake (μg NH 4 uptake/μg Chl a ) was greatest in

Nostoc muscorum after 48 h of exposure to 20,000 μg/L of lead. Compared to controls, the Pb test concentration 20,000 μg/L reduced a mmonium uptake by 72, 82, 61, and 26 % at 24, 48, 72, and 96 h, respectively (Rai and Raizada, 1989). Lead in combination with nickel at concentrations of 20,000 and 1,000 μg/L, respective ly, demonstrated an antagonistic effect on ammonium uptake. Compared to controls, ammonium uptake in the test group was reduced by 44.9, 54.1, 23.3, and 4% at 24, 48, 72, and 96 h, respectiv ely (Rai and Raizada, 1989). Lead in combination with chromium, both at concentrat ions of 20,000 μg/L, demonstrated a synergistic interaction with 24, 48, 72, and 96 h uptake rates reduced by 87.2, 88.5, 72.5, and 50 %, respectively, compared to controls (Rai and Raizada, 1989).AX7-181 Nostoc muscorum exposed to 20,000 μg Pb/L experien ced the greatest reduction in carbon fixation at 0.5 h of exposure: 62% compared to controls. Inhibiti on of carbon fixation in the test group was less pronounced after 1 and 2 h of expos ure: 29 and 13% of controls (Rai and Raizada, 1989). Lead in combination with nick el or chromium had synergistic effects to carbon

fixation. Lead and nickel concentrations of 20,000 and 1000 μg/L, respectively, resulted in 0.5, 1, and 2 h carbon fixation rates reduced by 93, 92, and 91%, respectively, compared to controls (Rai and Raizada, 1989). Lead with chromium at concentrations of 20,000 μg/L resulted in 0.5, 1, and 2 h carbon fixation rates reduced by 65, 58, a nd 50%, respectively, compared to controls. Nutrients such as nitrogen, phosphate, sodium acetate, sodium carbonate, and citric acid have been shown to protect against the toxic e ffects of Pb to algae (Jampani, 1988). Nitrogen compounds (ammonium chloride, potassium nitrate, sodium nitrate, sodium nitrite) protected Synechococcus aeruginosus from a lethal Pb-nitrate dose of 200,000 μg/L (Jampani, 1988). Two phosphates (K 2 HPO 4 and Na 2 HPO 4 ) were found to improve Synechococcus aeruginosus survival from 0 to 72% at 200,000 μg/L of Pb-nitrate (Jampani, 1988). Compared to controls, protein content wa s reduced by 54.2 and 51.9% in aquatic plants Vallisneria spiralis and Hydrilla verticillata , respectively, exposed to Pb for 7 days at 20,720 μg/L (Gupta and Chandra, 1994). Decreased soluble protein content has been observed

in Scenedesmus obliquus exposed to Pb-nitrate or Pb-acetate at concentrations greater than 30,000 μg/L, and in L. gibba at concentrations greater than 200,000 μg/L (Adam and Abdel-Basset, 1990; Miranda and Ilangovan, 1996). Lemna gibba also showed increased loss of soluble starch at concentrations >200,000 μg /L (Miranda and Ilangovan, 1996). Under the conditions described previously (Gupta and Chandra, 1994), EC 50 values for chlorophyll content were 14,504 and 18,648 μg/L for Vallisneria spiralis and Hydrilla verticillata , respectively (Gupta and Chandra, 1994). Effects to chlorophyll a content have been observed in Scenedesmus obliquus at Pb-nitrate and Pb-acetate con centrations >30,000 μg/L (Adam and Abdel-Basset, 1990).

Summary of Toxic Effects Observed in Single-Species Bioassays Algae and aquatic plants have a wide range in sensitivity to the effects of Pb in water. Both groups of primar y producers experience EC 50 values for growth inhibition between approximately 1000 and >100,000 μg/L (Jampani, 1988; Gaur et al., 1994; Bilgrami and Kumar,AX7-182 1997). The most sensitive primary producers reported in the literature for e ffects to growth were

Closterium acersoum and Azolla pinnata (Gaur et al., 1994; Bilgrami and Kumar, 1997). The least sensitive primary producers reported in the literature for effects to growth were Synechococcus aeruginosus and Lemna gibba (Jampani, 1988; Miranda and Ilangovan, 1996). Exposure to Pb in combination with other metals is generally less toxic to growth than exposure

to lead alone. Studies have shown that lead a dversely affects the metabolic processes of nitrate uptake, nitrogen fixation, ammonium uptake, a nd carbon fixation (Rai and Raizada, 1989). Lead

in combination with nickel or chromium produced synergisti c effects for nitrate uptake, nitrogenase activities, ammonium uptake, and carbon fixati on (Rai and Raizada, 1989).

Leads Effects on Primary Productivity

Lead nitrate and Pb-acetate have been s hown to have adverse effects on the primary productivity of aquatic plants in two water bodies in India (Jayar aj et al., 1992). One of the two

water bodies was a freshwater tank that receive s wastewater and supports a rich population of hyacinths, and the other was a wast ewater stabilization pond. Water quality characteristics in the freshwater tank were pH = 7.5, dissolved oxygen = 6 mg/L, and water hardness (CaCO 3 ) = 100 mg/L. Water quality characteristics in the wastewater pond were pH = 8.1, dissolved oxygen = 6.2 mg/L, and water hardness (CaCO 3 ) = 160 mg/L (Jayaraj et al., 1992). Lead nitrate concentrations of 500, 5000, 10,000, 25,000, and 50,000 μg/L were combined with appropriate water samples in light and dark bottles and susp ended in each of the water bodies for 4 h. The concentrations of Pb-acetate (5000, 1 0,000, 25,000, 50,000, and 100,000 μg/L) were applied in

the same manner. The EC 50 values were determined based on the concentra tion required to inhibit gross productivity (GP) and net productivity (NP) by 50% (Jayaraj et al., 1992). The

results demonstrated that Pb-nitrate was more toxic to primary production than Pb-acetate. In the freshwater tank, Pb-nitrate EC 50 values for GP and NP were 25,100 and 6310 μg/L, respectively, compared to Pb-acetate EC 50 values of 50,100 and 28,200 μg/L for GP and NP, respectively (Jayaraj et al., 1992). In the stabilization pond, Pb-nitrate EC 50 values for GP and NP were 31,600 and 28,200 μg/L, respectively, compared to Pb-acetate EC 50 values of 79,400 and 316 μg/L for GP and NP, respectively (Jayaraj et al., 1992). The higher toxicity reported in

the freshwater tank was attribut ed to differences in species composition and diversity. The freshwater tank was dominated by water hyacinths that decreased the photic zone available forAX7-183 photosynthesis and consumed a great deal of av ailable nutrients. The stabilization pond had a rich nutrient budget, resulti ng in improved alga growth a nd species diversity (Jayaraj et al., 1992).

AX7.2.4.3 Recent Studies on Effe cts of Lead on Consumers

This section focuses on the effects of Pb to a quatic biota including i nvertebrates, fish, and other biota with an aquatic life stage (e.g., amphibians). It is not intended to be a comprehensive

review of all research conducted. Rather, the intent is to illustrate the conc entrations and effects of Pb on freshwater and marine aquatic species. Eisler (2000) provides an overview of much of

the recent available literature on the toxicity of Pb to fish and aquatic invertebrates. An extensive literature search was conducted using nu merous electronic bibl iographic and database services (e.g., DIALOG, EPA ECOTOX) and limited temporally from 1986 to present. This temporal limit was due to the availability of the EPA water quality criteria report for the protection of aquatic life, released in 1986 (U.S. Environmental Protection Agency, 1986b).

Based on the results of the litera ture search and recent reviews of the toxicity of Pb (Eisler, 2000), numerous studies have been published on th e toxicity of Pb to aquatic consumers. Hardness, pH, temperature, and other factors ar e important considerations when characterizing

the acute and chronic to xicity of lead (Besser et al., 2005) (Section AX7.2.3.5). However, many of the studies reviewed did not report critical information on control mortality, water quality parameters, or statistical methods, making compar ing effects between stud ies difficult. Studies reporting only physiological responses to Pb exposure (e.g., reducti on of ALAD) are not discussed here, as this topic was covered more completely in Section AX7.2.3.4. This section provides a review of toxicity studies conducted with invertebra tes, fish, and other aquatic organisms.

Invertebrates Exposure of invertebrates to Pb can lead to adverse effects on reproduction, growth, survival, and metabolism (Eisler, 2000). The fo llowing presents informa tion on the toxicity of

Pb to invertebrates in fresh and marine waters.AX7-184 Freshwater Invertebrates Acute and chronic Pb toxicity data for fres hwater invertebrates are summarized in Table AX7-2.4.1. As described in Section AX7.2.3.5, water ha rdness is a critical factor governing the solubility, bioavailability, and ulti mately the toxicity of Pb. The acute and chronic toxicity of Pb increases with decreasing water hardness as Pb becomes more soluble and bioavailable to aquatic organisms. For example, Borgmann et al. (2005) examined the toxicity of 63 metals, including Pb, to Hyalella azteca at two levels of water hardness (soft water hardness, 18 mg

CaCO 3 /L; hard water, 124 mg CaCO 3 /L). Lead was 23 times more acutely toxic to H. azteca in

soft water than hard water. Besser et al. (2005) found that acute toxicity to H. azteca was also modified by water hardness. At a mean pH of 7.97 in soft water (hardness (CaCO 3 ) = 71 mg/L) mortality was >50%

for H. azteca at a dissolved Pb concentration of 151 μg/L . The LOEC for surv ival in hard water (hardness (CaCO 3 ) = 275 mg/L) at pH 8.27 was 192 μg/L as dissolved Pb and 466 μg/L as total Pb. Both waterborne and dietary Pb were found to contribute to reduced survival of H. azteca (Besser et al., 2005). Exposure duration may also play an important role in Pb toxicity in some species. For example, Kraak et al. (1994) reported that filtration in th e freshwater mussel Dreissena polymorpha was adversely affected at significantly lower Pb concentrations over 10 weeks of exposure than was the case after 48 h of exposure. The influence of pH on lead toxicity in freshwater invertebrates varies between invertebrate species. Over a 96- h exposure period, mortality increa sed with decreasing pH in the

bivalve Pisidium casertanum , while pH-independent mortality was reported for gastropod and crustacean species under similar exposure conditions (Mackie, 1989). Cladocerans ( Ceriodaphnia dubia ), amphipods ( H. azteca ), and mayflies ( Leptophlebia marginata ) were also more sensitive to Pb toxicity at lower pH levels (Schubauer-Berigan et al., 1993; Gerhardt, 1994). Lead was 100 times more toxic to the amphipod , Hyalella azteca , at a pH range of 5.0 to 6.0 (Mackie, 1989) than at a pH range of 7.0 to 8.5 (Schubauer-Berigan et al., 1993). The physiology of an aquatic organism at cer tain life stages may be important when determining the toxicity of metals to test orga nisms. For example, Bodar et al. (1989) exposed early life stages of Daphnia magna to concentrations of Pb(NO 3 ) 2 . The test medium had a pH

of 8.3 ∀ 0.2, water hardness (CaCO 3 ) of 150 mg/L, and temperature of 20 ∀ 1 Ε C. LeadAX7-185Table AX7-2.4.1. Effects of Lead to Freshwater and Marine InvertebratesSpecies ChemicalEndpoint:Conc. (μg/L)*Duration ofExposureWater Chemistry Test Type - Effect ReferenceFreshwaterCladoceran(Ceriodaphnia dubia)leadchlorideLC50:280>2,700>2,70048 h pH:6–6.57–7.58–8.5Hardness: 280-300 mg/L CaCO3static-survival Schubauer-Beriganet al. (1993)Worm(Lumbriculus variegatus)leadchlorideLC50:>8,000>8,000>8,00096 h pH:6–6.57–7.58–8.5Hardness: 280-300 mg/L CaCO3static-survival Schubauer-Beriganet al. (1993)Amphipod(Hyalella azteca)leadchlorideLC50:<90>5,400>5,40096 h pH:6–6.57–7.58–8.5Hardness: 280-300 mg/L CaCO3static-survival Schubauer-Beriganet al. (1993)Amphipod(Hyalella azteca)leadchlorideLC50:27 (20.1-36.4)8 days Hardness130 mg/LpH 7.8-8.6renewal, 1-weekold amphipodsMacLean et al.(1996)Amphipod(Hyalella azteca)leadchlorideLC50:60 (53.6-67.3)8 days Hardness130 mg/LpH 7.8–8.6renewal, 10- to 16-week oldamphipodsMacLean et al.(1996)Mayfly(Leptophlebia marginata)leadchlorideLC50:1090(400-133200)96 h pH:4.5acute - survival Gerhardt (1994)Mayfly(Leptophlebia marginata)leadchlorideLC50:500096 h pH7.0acute - survival Gerhardt (1994)AX7-186Table AX7-2.4.1 (cont’d). Effects of Leadto Freshwater and Marine InvertebratesSpecies ChemicalEndpoint:Conc. (μg/L)*Duration ofExposure Water Chemistry TestType - Effect ReferenceAmphipod(Hyalella azteca)leadnitrateLC50:10211896 h pH:5.05.56.0acute-survival Mackie (1989)Bivalve(Pisidium compressum)leadnitrateLC50:38,00021,30011,40096 h pH:3.54.04.5acute- survival Mackie (1989)Bivalve(Pisidium casertanum)leadnitrateLC50:23,60023,50056,00096 h pH:3.54.04.5acute- survival Mackie (1989)Gastropod(Amnicola limosa)leadnitrateLC50:10,30020,6009,50096 h pH:3.54.04.5acute- survival Mackie (1989)Mussel(Dreissena polymorpha)leadnitrateEC50:3709148 h10 wkspH = 7.9; Hardness = 150 mgCaCO3/L; Temp = 15ΕCrenewal - filtration Kraak et al. (1994)Mussel(Dreissena polymorpha)leadnitrateLT50:35872 days pH = 7.9; Hardness = 150 mgCaCO3/L; Temp = 15ΕCrenewal - filtration Kraak et al. (1994)Amphipod(Hyalella azteca)leadnitrateLC50:4.8 (3.3 - 7.1)7 days pH = 7.37 - 8.27Hardness = 18 mg CaCO3/LDOC = 0.28 mg/Lrenewal - survival Borgmann et al.(2005)Amphipod(Hyalella azteca)leadnitrateLC50:113 (101 -126)7 days pH = 8.21 - 8.46Hardness = 124 mg CaCO3/L;DOC = 1.1 mg/Lrenewal - survival Borgmann et al.(2005)AX7-187Table AX7-2.4.1 (cont’d). Effects of Leadto Freshwater and Marine InvertebratesSpecies ChemicalEndpoint:Conc. (Φg/L)*Duration ofExposureWater Chemistry Test Type - Effect ReferenceMayfly(Leptophlebia marginata)leadchlorideLC50:1090 (95% C.I. =133.2)>500096 h pH = 4.5 - 6.5;DOC - 21.6 mg Cl!1;Cond = 7.0ΦS cm!1renewal - survival Gerhardt (1994)Cladoceran(D. magna)leadnitrateLC50:0.4548 hpH = 8.3∀0.2Hardness (CaCO3) = 150 mg/LTemp= 20ΕCstatic -embryogenesisBodar et al. (1989)Cladoceran(D. magna)leadchlorideNOEC:26012 to 21 d Not specified renewal -reproductionEnserink et al.(1991)Cladoceran(D.magna)leadchlorideNOEC:27010 d Not specified renewal - growth Enserink et al.(1991)Amphipod(Hyalella azteca)lead LOEC:(Dissolved Pb)192(Total Pb) 46696 h pH = 8.27Hardness (CaCO3) = 275 mg/LTemp = 21.1ΕCflow through -survivalBesser et al. (2005)Tubificid worm(Tubifex tubifex)leadnitrateEC50:237 (183–316)142 (107–184)24 h48 hpH = 7.5–7.7Hardness = 245 mg/LTemp = 29.5–31ΕCstatic -immobilizationKhangarot (1991)MarineCopepod(Amphiascus tenuiremis)lead LC50: sediment2462 μgmetal/drysediment96 hpH = 7.7∀0.1Dissolved O2-6.3∀0.3 mg/LSalinity – 32 pptHagopian-Schlekatet al. (2001)Bivalve(Mytilus galloprovincialis)leadnitrateEC50: 221(58.9–346.3)LOEC : 50artificial seawater embryogenesis Beiras andAlbentosa (2003)* - Brackets after effect concentration are 95% confidence intervals.AX7-188 concentrations of <100 mg/L had no impact on Daphnia egg development. The authors suggested that this might due to the Daphnia egg structure, which consists of two layers: the inner vitelline layer and outer ch lorion layer. The chlorion layer in other species (e.g., rainbow trout) is known to adsorb metals, thereby, prev enting ionic injury to the developing embryo. Exposures to sediment-associated Pb can be toxic to sediment-dwelling organisms. In freshwater sediments, 48- h exposure of water fleas ( Daphnia magna ) to 7000 mg Pb/kg dw significantly reduced mobility, while exposur e to 13,400 mg Pb/kg dw for 24 h produced the same effect (Dave, 1992a,b). Longer-term (i.e., 14-day) exposure of midges ( Chironomus

tentans ) to sediments containing 31,900 mg Pb /kg dw resulted in 100% mortality.

Marine Invertebrates In estuarine environments, salinity is an im portant modifying factor to Pb toxicity. Verslycke et al. (2003) exposed the estuarine mysid Neomysis integer to individual metals, including Pb, and metal mixtures under ch anging salinity. Water temperature (20 ∀ 1 o C) and salinity were reported, although no other water qualit y parameters were available (e.g., pH, water hardness). At a salinity of 5 l , the reported LC 50 for Pb was 1140 μg/L (95% CI: 840, 1440 μg/L). At an increased salinity of 25 l , the toxicity of lead wa s substantially reduced (LC 50 = 4274 μg/L [3540, 5710 μg/L]) (Verslycke et al., 2003). Sensitivity to Pb can also vary between genders in some aquatic organisms. For example, Hagopian-Schlekat et al. (2001) examined the toxici ty of Pb-chloride in sediment and sediment pore water to female and male estuarine copepods Amphiascus tenuiremis . The reported LC 50 for total lead was 2462 mg Pb/kg dw (95% CI: 2097, 2891 mg Pb/kg dw). Gender effects were observed in that male copepods were more se nsitive (p = 0.038) to Pb than females as determined by generalized linear model analysis. Beiras and Albentosa (2003) examined th e inhibition of embryo development in

commercial bivalves Ruditapes decussatus and Mytilus galloprovincialis after exposure to concentrations of Pb(NO 3 ) 2 in seawater. No water chemistry parameters other than temperature were reported (test conducted at 20 o C). An EC 50 range for R. decussatus was reported as 156 to 312 μg/L, as insufficient data were available to calculate the actual EC 50 . The lowest observable effect concentration (LOEC) was 156 μg/L. For M. galloprovincialis , the EC 50 was 221 μg/L (95% CI: 58.9, 346.3) while the LOEC was reported as 50 μg/L.AX7-189 Fish The general symptoms of Pb t oxicity in fish include production of excess mucus, lordosis, anemia, darkening of the dorsal tail region, degeneration of the ca udal fin, destruction of spinal neurons, ALAD inhibition, growth inhibition, renal pathology, reproductive effects, growth inhibition, and mortality (Eisler, 2000). Toxicity in fish ha s been closely correlated with duration of Pb exposure and uptake (Eisler, 2000) . The following presents information on the toxicity of Pb to fish in fresh and marine wa ters. Table AX7-2.4.2 summar izes the effects of Pb

on freshwater and marine fish.

Freshwater Fish Many of the toxicity modifying factors described above and in Section AX7.2.3.5 (e.g., pH, DOC) for invertebrates are also important modifying factors for Pb toxicity to fish species. The effects of pH on Pb bioavailability and subs equent toxicity have b een well-studied (Sayer et al., 1989; Spry and Wiener, 1991; Schubauer-Be rigan et al., 1993; Stouthart et al., 1994; MacDonald et al., 2002; Rogers and Wood, 2003). Schubauer-Ber igan et al. (1993) exposed

fathead minnow to Pb-chlorid e over 96 hours. The reported LC 50 ranged from 810 to >5400 μg/L at pH 6 to 6.5 and pH 7 to 8.5, respectively. Water hardness also has a strong influence on the effects of lead to fish. Chronic exposure of rainbow trout fry to Pb in soft wate r resulted in spinal deformities at 71 to 146 μg/L after 2 months of exposure (Sauter et al., 1976) or 13.2 to 27 μg/L (Davies and Everhart, 1973; Davies et al., 1976), after 19 months of exposure. When exposed to Pb in hard water, only 0 and 10% of the trout ( Oncorhynchus mykiss ) developed spinal deformities at measured Pb concentrations of 190 and 380 μg/L, respectively. In soft water, 44 and 97% of the trout developed spinal deformities at concentrations of 31 and 62 μg/L, respectively (Davies et al., 1976). The maximum acceptable toxicant concentra tion (MATC) for rainbow trout fry in soft water was 4.1 to 7.6 μg/L (Davies et al., 1976) , while the MATC for brook trout was 58 to 119 μg/L (Holcombe et al., 1976). Histological re productive abnormalities were noted in mature male rainbow trout at 10 μg/L Pb-nitrate (Ruby et al., 1993). Schwartz et al. (2004) examin ed the influence of NOM on Pb toxicity to rainbow trout exposed for 96 h in a static system. The pH of the exposure system ranged between 6.5 and 7.0, temperature was maintained between 9 and 11 Ε C, and Pb was added as PbCl 2 . NOM from aAX7-190Table AX7-2.4.2. Effects of Pb toFreshwater and Marine FishSpecies ChemicalEndpoint:Conc. (Φg/L)Duration ofExposureWater Chemistry Comments ReferenceFreshwaterFathead minnow(Pimephales promelas)leadchlorideLC50:810>5,400>5,40096 h pH:6–6.57–7.58–8.5Hardness: 280-300mg/L CaCO3static, measured Schubauer-Beriganet al. (1993)Rainbow trout - maturemales(Oncorhynchus mykiss)leadnitrateReproductive effects:1012 days Hardness 128 mg/LCaCO3DecreasedspermatocytedevelopmentRuby et al. (1993)Fathead minnow(Pimephales promelas)leadacetateReproductive Effects:50029 days pH:7.5–8.5;Hardness 130 mg/LCaCO3;22–25ΕC(Pb 95% soluble)Fewer viable eggsproduced, testiculardamageWeber (1993)Rainbow trout – Juvenile(Oncorhynchus mykiss)leadnitrateLC50: 1000(800 - 1400)96 h pH:7.9–8.0DOC = 3 mg/LHardness (CaCO3) =140 mg/LFlow through - Survival Rogers and Wood(2003)Common carp(Cyprinus carpio)notreportedLC50:6.5 cm fish – 10303.5 cm fish – 30096 h pH:7.1Temperature–15ΕCOxygen sat. 6.4 mg/Lstatic-renewal -SurvivalAlam and Maughan(1995)AX7-191 number of U.S. rivers and lakes was th en added to the test system, and the LT 50 was reported.

NOM was found to reduce the toxic eff ects of Pb to rainbow trout. Fish size is an important variable in determ ining the adverse effects of Pb. Alam and Maughan (1995) exposed two different sizes of common carp ( Cyprinus carpio ) to Pb concentrations to observed effects on carp mortal ity. Water chemistry parameters were reported (pH = 7.1; temperature = 20 Ε C). Smaller fish (3.5 cm) were found to be more sensitive to Pb than were larger fish (6.5 cm). The reported LC 50 s were 0.44 mg/L and 1.03 mg/L, respectively.

Marine Fish

There were no studies available that examined th e toxicity of Pb to marine fish species for the time period examined (1986 to present). Howe ver, Eisler (2000) review ed available research on Pb toxicity to marine species and reported studies done prior to 1986. Acute toxicity values ranged from 50 μg/L to 300,000 μg/L in plaice ( Pleuronectes platessa ) exposed to organic and inorganic forms of Pb (Eisler, 2000). Organol ead compounds (e.g., tetramet hyl Pb, tetraethyl Pb, triethyl Pb, diethyl Pb) were generally more t oxic to plaice than inorganic Pb (Maddock and Taylor, 1980).

Other Aquatic Biota A paucity of data exist on the effects of Pb to growth, reproduc tion, and survival of aquatic stages of frogs and turtles. Rice et al. (1999) exposed frog larvae ( Rana catesbeiana ) to 780 μg Pb/L and two oxygen concentrations (3.5 or 7.85 mg/L) for 7 days (Table AX7-2.4.3). Exposure conditions included wate r hardness of 233 to 244 mg CaCO 3 /L, pH from 7.85 to 7.9, and temperature at 23 Ε C. Frog larvae were found to displa y little to no activ ity in the low oxygen and high Pb treatment. Hypoxia-like behavi or was exhibited in larvae exposed to both low and high oxygen concentrations and high Pb. Therefore, larvae of R. catesbeiana showed sensitivity to Pb and responded with hypoxia-like behavior. Additionally, the larvae in the Pb

treatment were found to have lost body mass relative to controls a nd the other treatments. Rice et al. (1999) suggested that the decrease in mass likely indicated the beginning of a period of reduced growth rate. Larvae exposed for l onger periods (>4 weeks) were smaller and metamorphosed later compared to unexposed individuals.AX7-192Table AX7-2.4.3. Nonlethal Effects in AmphibiansSpecies ChemicalEndpoint:ConcentrationDuration ofExposureWater Chemistry Comments ReferenceFrogs(Rana ridibunda)lead nitrate Biochemical effects:14,000 μg/L30 days Not specified Hepatic ALAD decreased by90%Vogiatzis andLoumbourdis (1999)Frogs(Bufo arenarum)Mortality:16 mg Pb2+/L5 days Not specified Effects reported include erraticswimming, loss of equilibriumHerkovits and Perez-Coll(1991)Frogs(Rana catesbeiana)Hypoxia-like behavior:780 μg/L7 days O2= 3.5-7.85 mg/LpH = 7.85-7.9Temp = 23ΕCCaCO3= 233-244mg/LLarvae used Rice et al. (1999)Turtle Hatchlings(Trachemys scripta)lead acetate NOEL: 100 μg/g(Survival andbehavior)4 weeks N/A Exposure via single injection Burger et al. (1998)AX7-193 Herkovits and Pérez-Coll (1991) examined Pb toxicity to amphibian larvae ( Bufo

arenarum ). Larvae (n = 50) were exposed for up to 120 h at two Pb concentrations, 8 mg Pb 2+ /L and 16 mg Pb 2+ /L. Relative to controls, the 8 mg Pb 2+ /L treatment group exhibited 40% mortality and the 16 mg Pb 2+ /L group 60% mortality after 120 h (p < 0.05). The authors reported behavioral effects, erratic swim ming, and loss of equilibrium during the tests, symptoms that are

consistent with the action of Pb on the central and peripheral nervous systems (Rice et al., 1999). Behavior (i.e., righting, body tur nover, seeking cover), growth, and survival of hatchling slider turtles ( Trachemys scripta ) exposed to Pb-acetate were i nvestigated in one study (Burger et al., 1998). In the fi rst part of the study, 6-month-old ha tchlings received single Pb-acetate injections at 50 or 100 μg/g body weight (bw). In the second part of the study, 3-week-old turtles were injected once with doses of 250, 1000 or 2500 μg/g bw. There were no differences in survival, growth, or behavior for hatchlings in the first st udy, however, several effects were reported from the second part of the study at dos es in the range of 250 to 2,500 μg/g bw. As the dose increased, so did the plastr on length (i.e., ventral section of the shell), carapace length, and weight. The highest dose group had the lowest survival rate with an LD 50 of 500 μg/g bw.

Behavioral effects included slower times of ri ghting behavior and seeki ng cover. The authors suggested a NOEL of 100 μg/g bw for slid er turtles for survival and behavior.

AX7.2.4.4 Recent Studies on Effect s of Lead on Decomposers In this section, decomposers are defined as being bacteria and other microorganisms. Many invertebrates are also poten tially considered decomposers, but the effects of Pb to

invertebrates have been described in previous sections. There we re no toxicity studies located on the effects of Pb to aquatic decomposers in the time period of interest.

AX7.2.4.5 Summary

Lead in all its forms is known to cause adverse effects in aquatic organisms (Eisler, 2000). Effects to algal growth have been observ ed at b concentrations ranging from 100 to 200,000 μg/L. Clinical signs of Pb toxicity in pl ants include the deformation and disintegration of algae cells and a shortened exponential growth phase. Other effects of Pb include a blocking of the pathways that lead to pigment synthesi s, thus affecting photos ynthesis, cell cycle and division, and ultimately resulting in death. The toxicity of Pb to macrophyte growth has beenAX7-194 studied using Spirodela polyrhiza , Azolla pinnata , and Lemna gibba . Test durations ranged from 4 to 25 days and test concentrations ranged between 49.7 and 500,000 μg/L. Waterborne Pb is highly toxic to aquatic organisms, with toxi city varying with the species and life stage tested, duration of ex posure, form of Pb tested, and water quality characteristics. Among the species tested, aquatic invertebrates, such as amphipods and water fleas, were the most sensitive to the effects of Pb, with adve rse effects being reported as low as 0.45 μg/L. Effects concentrations for aquatic invertebra tes were found to range from 0.45 to 8000 μg/L.

Freshwater fish demonstrated adverse effects at concentrations ranging from 10 to >5400 μg/L, depending generally upon water qual ity parameters. Amphibians tend to be relatively Pb tolerant (Eisler, 2000); however, they may exhibit decrea sed enzyme activity (e.g., ALAD reduction) and changes in behavior (e.g., hypoxia response behavior). Lead tends to be more toxic with longerterm exposures. The primary focus of this section was on eff ects at the individual level of organization. This narrow focus is primarily due to a lack of information on the effects of lead on higher levels of organization (e.g., populations, communities, and ecosystems). The impact of lead at higher levels of biological organization should be cons idered in future reviews of lead toxicity. In considering these higher levels of organiza tion, cause-effect models and relationships that examine the roles of life history strategy and optimal foraging theory, community processes and associated theory, and ecosystem processes and th e influences of redundanc y/interactions should

be taken into account.



### AX7.2.5 Effects of Lead on Natural Aquatic Ecosystems

Introduction As discussed previously, lead exposure may adve rsely affect organisms at different levels of organization, i.e., individual organisms, popula tions, communities, or ecosystems. Generally, however, there is insufficient information availabl e for single materials in controlled studies to permit evaluation of specific impacts on higher levels of organizati on (beyond the individual organism). Potential effects at the population leve l or higher are, of necessity, extrapolated from

individual level studies. Av ailable population, community, or ecosystem level studies are typically conducted at sites that have been c ontaminated or adversely affected by multiple stressors (several chemicals alone or combined with physical or bi ological stressors). Therefore,AX7-195 the best documented links between lead and e ffects on the environment are with effects on individual organisms.

This section discusses the e ffects of Pb on natural aquati c ecosystems. Such effects include changes in species composition and richne ss, ecosystem function, and energy flow due to Pb stress. The format of this section genera lly follows a conceptual framework for discussing the effects of a stressor such as Pb on an ecosy stem. This conceptual framework was developed by the EPA Science Advisory Board (Young and Sa nzone, 2002). The essential attributes used to describe ecological condition include lands cape condition, biotic condition, chemical and physical characteristics, ecological proce sses, hydrology and geomorphology and natural disturbance regimes. The majority of the pub lished literature pertaining to Pb and aquatic ecosystems focuses on the biotic co ndition, one of several essential a ttributes of an ecosystem as described in Young and Sanzone (2002). For the biotic condition, the SAB framework identifies community extent, community composition, tr ophic structure, community dynamics, and physical structure as factors for assessing ecosy stem health. Other factors for assessing the

biotic condition such as effect s of Pb on organs, species, popula tions, and organism conditions (e.g., physiological status) were discu ssed in Sections AX7.2.3 and AX7.2.4. For natural aquatic ecosystems, the focus of study in the ge neral literature has been on evaluating ecological stress where th e sources of Pb were from ur ban and mining effluents rather than atmospheric deposition (Poulton et al., 1995; Deacon et al., 2001; Mucha et al., 2003). The atmospheric deposition of Pb in remote lakes has been evaluated; however, the direct effects of Pb on aquatic ecosystems was not evaluated in many cases (Larsen, 1983; Köck et al., 1996; Allen-Gil et al., 1997; Outridge, 1999; Lotter et al., 2002). In othe r studies, although Pb deposition was studied, the effect s of acid deposition on aquatic life were th e focus of the study and perceived to be more relevant (Mannio, 2001; Nyberg et al., 2001). Finally, the effects of Pb, other metals, and acidification on phytoplan kton have only been inferred based on the paleolimnolical record (Tolonen and Jaakkola, 1983; Rybak et al., 1989 ). The statistical methods used when evaluating the effects of Pb on aquatic ecosystems are important, as more than one variable may be related to the obser ved effect. Studied va riables include water hardness, pH, temperature, and physical factors such as embeddedness, dominant substrate, and velocity. In most cases, single va riable statistical techniques were used to evaluate the data.AX7-196 However, in other cases multivariate techniques we re used. Therefore, where appropriate, some

detail on the statistical met hods used is presented. Although most of the available studies discus sed in this section focus on the biotic condition, one case study examining multiple com ponents of the EPA conceptual framework is

also included. The remainder of th is section describes the effects of Pb on the biotic condition.

AX7.2.5.1 Case Study: Coeur d’Alene River Watershed

The Coeur d’Alene River watershed is an area of Idaho impacted by Pb and other metals from historic mining waste releases. Mare t et al. (2003) examined several ecological components to determine any negative associations with metals and the watershed communities. The variables examined and associated ecologi cal conditions are presented in Table AX7-2.5.1.

In addition to measurements of non-metal va riables (e.g., dissolved oxygen levels, water temperature and pH, embeddedness), Cd, Pb, and Zn le vels were also compared in affected sites

versus reference sites.

Some of the above non-metal variables are im portant to macroinvertebrate communities. For example, a stream with highly embedded subs trate can have a lower number of individuals within a species or a different species compos ition compared to a stream with less embeddedness (Waters, 1995). Macroinvertebrates from the Ephe meroptera (mayflies), Pl ecoptera (stoneflies), and Trichoptera (caddisflies) (EPT ) group inhabit the surf ace of cobble and the interstitial spaces

between and underneath cobble. When substrate is embedded, these inters titial spaces are filled, leaving less habitat space for EP T taxa. In another example, water temperature is important; some macroinvertebrates (e.g., ston eflies) are usually only found in cooler water (Harper and Stewart, 1984). Of the variables examined only metal concen trations, mine densit y, site elevation, and water temperature were significan tly different between reference and mine-affected sites. A Mann-Whitney t-test was used to evaluate statistical differences between reference and test sites for physical and water quality parameters , while Spearman’s rank correlation matrices were used to compare all possible response and expl anatory variables. Lead concentrations were significantly correlated with th e number of mines in proxim ity to the watershed. Lead concentrations in sediment and water were strongly correlated to Pb leve ls in whole caddisflies,

r 2 = 0.90 and 0.63, respectively. Furthermore, mine de nsity was significantly correlated to Pb inAX7-197 Table AX7-2.5.1. Ecological Attributed Studies by Maret et al. (2003) in the Coeur d’Alene Watershed

Ecological Attribute Subcategory Measure Landscape condition Areal extent landscape pattern Basin area (km 2 ) Production mine density/km 2

Biotic Condition Organism condition population structure/ dynamics Caddisfly tissue concentrations (mg/kg) Number of EPT taxa Density of EPT individuals (no./m 2 )

Chemical/physical characteristics Chemical/physical parameters Dissolved oxygen (mg/L) Specific conductance (μS/cm) Water temperature (E Ε C) pH Water hardness (mg/L) Total NO 3 (mg/L) Total P (mg/L) Dissolved NH 3 (mg/L) Sediment Cd, Pb, Zn (mg/kg) Dissolved Cd, Pb, Zn in water (mg/L)

Ecological processes — None measured

Hydrology/geomorphology Channel morphology and distribution Site elevation (m) Stream gradient (%) Stream discharge (m 3 /s) Stream width (m) Stream depth (m) Open canopy (%) Stream velocity (m/s) Embeddedness (%) Dominant substrate (mm)

Natural disturbance regimes — None measured

tissue, r 2 = 0.64. Although temperature was signifi cantly different between reference and mine-affected sites, temperature conditions were concluded to be non-lim iting to aquatic life. For example, reference and mine-affected sites ha d at least 15 and 13 obligate cold-water taxa, respectively. A significant negative correlation between Pb in the water column (0.5 to 30 μg/L dissolved) and total taxa richness , EPT taxa richness, and the num ber of metal-sensitive mayflyAX7-198 species was observed. Similar, significant negati ve correlations were found between sediment Pb levels (132 to 6252 μg/g) and the same macr oinvertebrate community metrics and caddisfly tissue levels. Negative correla tions were also found between Cd and Zn in the water and sediment and the macroinvertebrate community metr ics. In an analysis of cumulative toxicity, Pb was judged to be the most significant metal in sediment related to the cumulative toxicity measured. This study provided multiple lines of evidence (i.e., mine density, metal concentrations, bioaccumulation in caddisfly ti ssue and benthic invertebrate assemblage structure) of the negative impacts of mining in the Coeur d’Alene River, suggesting that Pb (and other metals) were primary contributors to th e effects observed in the Coeur d’Alene River watershed (Maret et al., 2003).

AX7.2.5.2 Biotic Condition In an evaluation of the biotic condition, the SAB framework described by Young and Sanzone (2002) identifies co mmunity extent, community co mposition, trophic structure, community dynamics, and physical structure as essential ecological at tributes for assessing ecosystem health. The following two sections describe the effects of Pb on community composition, community dynamics, and trophic struct ure. To date, no available studies were located on the effects of Pb on physical struct ure (e.g., change in ripa rian tree canopy height, ecosystem succession).

Ecosystems and Communities, Community Composition To measure community composition, an invent ory of the species/taxa found in the ecological system must be conducted. Accord ing the SAB framework, useful measures of composition include the total number of species or taxonomic units, their relative abundance, presence and abundance of native and non-native species, and information on the presence and abundance of focal or special interest speci es (Young and Sanzone, 2002). Focal or special species of interest can be those that play a critic al role in ecosystem processes such as flows of materials or energy within complex food- webs (Young and Sanzone, 2002). Community composition as assessed in Pb studies has included the following measures.AX7-199 • Changes in energy flow or nutrient cycling:

o Increased or decreased respiration or biomass

o Increased or decreased turnover/cycling of nutrients

• Changes in community structure:

o Reduced species abundance (i.e., the total num ber of individuals of a species within a given area or community)

o Reduced species richness (i.e., the numb er of different species present in a community)

o Reduced species diversity (i.e., a measur e of both species abundance and species richness)
Investigators have evaluated the effects of Pb on aquatic communities through microcosm and mesocosm studies in natural aquatic systems. Field studies in the ge neral literature have focused on natural systems that were affected by metal stress from various anthropogenic sources. In most of those natural systems, the sources evaluated were from direct mining waste inputs, rather than atmospheric deposition, of Pb. Studies published since the 1986 Lead AQCD (U.S. Environmental Protection Agency, 1986a) that describe the effects of Pb on natural aquatic ecosystems are presented below and summarized in Table AX7-2.5.2. Studies included here evaluated the effects of Pb on watersheds, landscapes, aquatic ecosystems, aquatic communities, biodiversity, lakes, rivers, streams, estu aries, wetlands, and species interaction.

Aquatic Microcosm Studies The examination of simulated aquatic ecosy stems (i.e., microcosms) provides limited information on the effects of pollutants on natura l systems. Microcosm studies typically focus on only a few aspects of the natural system a nd do not incorporate a ll of the ecological, chemical, or biological interactions. Nevert heless, a few microcos m studies have been conducted that indicate potential e ffects of Pb on the community structure of aquatic ecosystems.

Fernandez-Leborans and Antonio- García (1988) evaluated the effect of Pb on a natural community of freshwater protozoans in simula ted aquatic ecosystems and found a reduction in the abundance and composition of protozoan specie s with increasing Pb concentrations (0.05 to 1.0 mg/L) compared to controls. Studies with marine protozoan communities in laboratory microcosms indicated that waterborne Pb expos ure reduced protozoan abundance, biomass, andAX7-200Table AX7-2.5.2. Essential Ecological Attributes for Natural Aquatic Ecosystems Affected by LeadCategory SpeciesConditionMeasuresExposureMediumLocationExposureConcentrationsOtherMetalsPresentReferenceBiotic ConditionEcosystems andCommunities-CommunityCompositionProtozoan community Reduced speciesabundance anddiversityMarine water Laboratory microcosm 0.02– 0.05 mg/L N Fernandez-Leborans andNovillo (1992)Protozoan community Reduced speciesabundanceFreshwaterwaterLaboratory microcosm 0.05–1 mg/L N Fernandez-Leborans andAntonio-García(1988)Protist community Reduced speciesabundance anddiversityMarine water Laboratory microcosm 1 mg/L N Fernandez-Leborans andNovillo (1994)Meiofauna community Reduced abundance MarinesedimentLaboratory microcosm 177 mg/kg dw Y Millward et al.(2001)Algal community Increased respiration Freshwater Domestic waterstabilization pond25–80 mg/L ? Jayaraj et al. (1992)Algal community Decreased primaryproductivityFreshwater Sharana BasaveshwaraTank, India6–32 mg/L ? Jayaraj et al. (1992)Meiobenthiccommunity(primarily nematodes)Reduced speciesabundanceNo effect onabundanceMarinesedimentLaboratory microcosm 1343 mg/kg dw1580 mg/kg dwN Austen andMcEvoy (1997)AX7-201Table AX7-2.5.2 (cont’d). Essential Ecological Attributesfor Natural Aquatic Ecosystems Affected by LeadCategory SpeciesConditionMeasuresExposureMedium LocationExposureConcentrationsOtherMetalsPresent ReferenceMacroinvertebratecommunityLower totalabundance, decreasedtaxa, and EPTrichness, largerpercentage of tolerantspecies of benthicmacroinvertebrates.Freshwaterand sedimentMining sites in theUpper Colorado Basin<0.001–0.02 mg/L145-850 mg/kg dw(<63 μM fraction)Y Deacon et al.(2001); Mize andDeacon (2002)MacroinvertebratecommunityNegatively correlatedwith species richnessand diversity indicesEstuarysedimentDouro Estuary, Portugal 0.25–192 mg/kg dw Y Mucha et al. (2003)MacroinvertebratecommunityReduced speciesabundanceFreshwatersedimentRiver Ill and tributaries,France1–16 mg/kg dw Y Rosso et al. (1994)Fish, crustacean andmacroinvertebratecommunityCorrelation withchanges in speciesabundance anddistributionMarineSedimentSpencer Gulf, SouthAustralia156–5270 mg/kgdwY Ward and Young(1982);Ward andHutchings (1996)Chironomid community Reduced chironomidrichnessWholeorganismresidueNew Brunswick, Canada 40.3–1,387 mg/kgdw (periphyton)1.6–131 mg/kg dw(chironomid tissue)Y Swansburg et al.(2002)MacroinvertebratecommunityLead in tissuesnegatively correlatedwith taxa richness,EPT richness,chironomid richness,and species density.WholeorganismresidueClark Fork River, MT 32.2–67.1 mg/kgdwY Poulton et al.(1995)MacroinvertebratecommunityLead in tissuesnegatively correlatedwith EPT richnessand abundance.BiofilmresiduesBoulder River, MT 32–1540 mg/kg dw Y Rhea et al. (2004)AX7-202Table AX7-2.5.2 (cont’d). Essential Ecological Attributesfor Natural Aquatic Ecosystems Affected by LeadCategorySpeciesCondition MeasuresExposureMediumLocationExposureConcentrationsOtherMetalsPresentReferenceMacroinvertebrateCommunityLead in tissues andsediment not correlatedto diversity andrichnessSedimentand wholeorganismresidueAquashicola Creektributaries, Palmerton,PA7.5–59.5 mg/kg dw(sediment)0.25-6.03 mg/kg dw(macroinvertebrates)Y Carline and Jobsis(1993)Fish Community Lead in tissues andsediment not correlatedto diversity andrichnessSedimentand wholeorganismresidueAquashicola Creektributaries, Palmerton,PA7.5–59.5 mg/kg dw(sediment)0.1-0.86 mg/kg dw(fish)Y Carline and Jobsis(1993)Ecosystems andCommunities-CommunityDynamics andTrophic StructureSnails and tadpoles Lead affected predator-prey interactionsSediment Outdoor mini-ecosystemsNot cited Y Lefcort et al. (1999)Snails and caddisflies No avoidance ofpredator by snail.Caddisfly did respondto predatorWater Field microcosm forsnail; in-streamdisturbance forcaddisfly27.7–277.6 mg/kgdw (snail tissue)223–13,507 mg/kgdw (caddisfly tissue)Y Lefcort et al. (2000)Fathead minnow Feeding behavioralteredWater Laboratory microcosm0.5–1.0 mg/L N Weber (1996)American toad No avoidance of lead Water Laboratory microcosm 0.5–1.0 mg/LN Steele et al. (1991)Mummichog Feeding behavioraltered and predatoravoidance affectedWater Laboratory 0.3–1.0 mg/L N Weis and Weis(1998)AX7-203 diversity at concentrations of 0.02 to 1.0 mg/L Pb (Fernandez-Leborans and Novillo, 1992, 1994). Austen and McEvoy (1997) studied the eff ects of Pb on an estuarine meiobenthic community (mainly nematodes) in a microcos m setting using sediment samples collected offshore from England. A multivariate analysis ofsimilarities (ANOSIM) test with square roottransformed data was used to evaluate differen ces between treatments and controls. Lead was found to significantly affect species abundance at 1343 mg/kg dw relative to a control at 56 mg/kg dw, but no significant adverse effects were observed at the highest dose tested, 1580 mg/kg dw. The authors did not attempt to explain why the 1580 mg/kg dw dose was not significant while the 1343 mg/kg dw dose was. N one of the Pb exposures were significantly different than the controls based on separate univariate tests of abundance, richness, and diversity. There were no other confounding metals in the Pb tests, as the experiments were with a single metal dose. In one other mesocosm st udy, the effects of a mixture of metals (Cu, Cd, Pb, Hg, and Zn) on a salt marsh meiofaunal comm unity were evaluated (Millward et al., 2001). After 30-days exposure, significant reduc tions in copepod, gastropod, and bivalve abundances were observed at the highest Pb e xposure concentration, 177 mg/kg dw. Ostracods

and nematodes were not affected. The authors be lieved that the response of the meiofauna taxa to metals was in part due to the various feedi ng strategies in that de posit feeders were most

affected.

Natural Aquatic Ecosystem Studies Lead stress in aquatic ecosystems has also been evaluated in natural communities. Studies examining community-sca le endpoints, however, are comp lex, and interpretation can be confounded by the variability found in natural syst ems and the presence of multiple stressors. Natural systems frequently contain multiple metals, making it difficult to attribute observed adverse effects to single metals. For example, macroinvertebrate communities have been widely studied with respect to metals contaminati on and community composition and species richness (Winner et al., 1980; Chadwick et al., 1986; Clements, 1994). In these studies, multiple metals

are evaluated and correlations between observed community level effects are ascertained. The results often indicate a correlation between the pr esence of one or more metals (or total metals) and the negative effects observed. While, corr elation may imply a relationship between twoAX7-204 variables, it does not imply causa tion of effects. The following studies suggest an association

between Pb concentration and an alteration of community structure and function (see summary in Table AX7-2.6.2):

Reduced Primary Productivity and Respiration Jayaraj et al. (1992) examined the effects of Pb on primary productivity and respiration in an algal community of two water bodies. Concentr ations of Pb in water (6 to 80 mg/L) were found to significantly reduce primary productivity and increase respiration. The authors suggested that increased respirat ion indicated a greater tolerance to or adaptive mechanisms of the resident heterotrophs to cope with lead stress.

Alterations of Community Structure Deacon et al. (2001) studied a macroinvertebra te community in mine-affected waters of Colorado. Initially, transplant ed bryophytes were used to assess whether metals could

bioaccumulate at various mine-affected and unaffected sites (Deacon et al., 2001; Mize and Deacon, 2002). Lead was bioaccumulated by the bryophytes, and median tissue concentrations at mine-affected sites (34 to 299 μ g/g dw) were higher than at reference sites (2.5 to 14.7 μ g/g dw). Lead concentrations in surface water and sediment ranged from <0.001 to 0.02 mg/L and 145 to 850 mg/kg dw (<63 μm fraction), respectively. The same sites were also evaluated for the

effects of various metals on macroinvertebrate communities. Values of total abundance, taxa richness, mayfly, and stonefly abundance were redu ced at mining sites. Lead levels along with Cd, Cu, and Zn were correlated with re duced abundance and diversity indices. Macrobenthic communities studied in a Portuguese estuary were affected by Pb at a range from 0.25 to 192 mg/kg dw (Mucha et al., 2003). Species richness was decr eased in areas with

increased Pb concentrations in the sediment. Interpretation of Pb effects was complicated by other non-metal stressors, namely sediment pa rticle size and organic matter content. Furthermore, other metals were present (e.g., Al , Cu, Cr, Mn, Zn) and may have affected the community (Mucha et al., 2003). The effects of Pb on oligochaetes in the I ll River and its tributaries in France were evaluated by Rosso et al. (1994). Lead in sediment (5 to 16 μ g/g dw at affected sites) was positively correlated to the abundance of the oligochaete, Nais sp., and negatively correlated toAX7-205 Tubificidae abundance. Lead was the only metal that was positively correlated to Nais species, while other metals were negatively correlate d to Tubificidae (Ro sso et al., 1994). The effects of metals and particle size on structuring epibenthic sea grass fauna (fish, mollusks, crustaceans, and polychaetes) was eval uated near a Pb smelter in South Australia (Ward and Young, 1982; Ward and Hutchings, 1996). Effluent from the smelter was the primary source of Pb and other metal contamination. Sp ecies richness and composition were evaluated near the Pb smelter along with metal concentrations in sediment. Lead levels in sediment (up to 5270 mg/kg dw) correlated with ne gative effects on species ric hness and composition, while the

other metals evaluated had similar correlations. Therefore, Pb alone could not be identified as the sole metal causing stress.

Tissue Bioaccumulation Associated with Alterations of Community Structure

Several studies have examined the bioaccumu lation of lead in a quatic systems with indices of community structure and function. A focused study on changes in Chironomidae community composition in relation to metal mine s (New Brunswick, Canada) identified changes in Chironomidae richness (Swansburg et al., 2002) . Lead was not detected (detection limit not given for any matrix) in the wate r column at any site. However, Pb levels in periphyton were significantly higher at mining sites (40.3 to 1387 mg /kg dw) compared to reference sites (not detected [ND], 33.3 mg/kg dw). Furthermore, Pb in chironomids was significantly higher at mine-affected sites (1.6 to 131 mg/kg dw) compar ed to reference sites (ND, 10.2 mg/kg dw).

The concentrations in biota indicate that Pb is mobile and available to the aquatic community even though water concentrations were undetectab le. Chironomidae richness was reduced at the sites receiving mining effluent c ontaining Pb, Cd, Cu, and Zn. In another study, macroinvertebra te lead tissue concentrations (32.2 to 67.1 mg/kg dw at affected sites) collected from the Clark Fork River, Montana correlated negatively with total richness, EPT richness, and dens ity (Poulton et al., 1995). Mean Pb levels were as high as 67.1 mg/kg dw at sites most affected by lead. However, other metals, including Cd, Cu, and Zn, also were negatively correlated with total richness and EPT richness. Therefore, attribution of the observed effects to Pb is difficult, as other metals may be contributing factors. In Montana, the potential effects of metals on macroinvertebrate communities in the Boulder River watershed were evaluated (Rhea et al., 2004). Similar to the approach taken byAX7-206 Poulton et al. (1995), the effects on richness and abundance of EPT taxa were compared to metal concentrations in tissue (i.e., biofilm and macroi nvertebrates). Lead levels in biofilm (32 to 1540 mg/kg dw) were significantly correlated with habitat scores and macroinvertebrate indices (e.g., EPT taxa). However, macroinvertebrate tiss ue Pb levels were not significantly correlated with macroinvertebrate community level metrics. As with most natural systems with potential mine impacts, other metals also correlated with community level effects. However, the authors

indicated that Pb concentrations in biofilm appeared to have the most significant impact on

macroinvertebrate metrics. A detailed investigation of sediment, macr oinvertebrates, and fish was conducted for tributaries in the Aquashicola Creek watershed near a former zinc smelter in Palmerton, PA (Carline and Jobsis, 1993). The smelter deposit ed large amounts of Cd, Cu, Pb, and Zn on the surrounding landscape during its operation from 1898 to 1980. The goal of the study was to

evaluate if there was a trend in the metal levels in sediment, macroinvertebrate and fish tissue, and community indices going away from the sme lter. Sites were chosen, from 7.8 to 24.6 km

from the smelter. There were no clear associat ions between proximity to the smelter and Pb levels in sediment, macroinvertebrate tissue, and fish tissue. Furthermore, there were no associations between proximity to the smelter and macroinvertebrate and fish diversity and richness. The authors suggested th at the transport of metals in the watershed has decreased since the smelter ceased operation, and thereby no effects were observed.

Ecosystems and Communities, Community Dynamics, and Trophic Structure As described in the SAB framework, community dynamics include interspecies interactions such as competition, pred ation, and succession (Young and Sanzone, 2002).

Measures of biotic interacti ons (e.g., levels of seed disper sal, prevalence of disease in populations of focal species) provide important information about community condition. If the community dynamics are disrupted, then the trophic structure may also be disrupted. According to the SAB framework, trophic structure refers to the distribution of sp ecies/taxa and functional groups across trophic levels. Measures of trophic structure include food web complexity and the presence/absence of top predators or dominant herbivores. Theref ore, this section discusses how aquatic species interactions can be affected by Pb. Examples of species interactions can include:AX7-207 • Predator-prey interactions (e.g., re duced avoidance of predators)

• Prey consumption rate (e.g., incr ease or decrease in feeding)

• Species competition (e.g., interference with another species, increased aggressive behavior)

• Species tolerance/sensitivity (e.g., the emergence of a dominant species due to contaminant tolerance or sensitivity)
Species interactions are highly relevant to a discussion about the effects of Pb on natural aquatic ecosystems, because effects on species interactions could potentially affect ecosystem function and diversity. Some examples of Pb induced changes in species interactions are presented below (see summary in Table AX7-2.5.2).

Predator-prey Interactions Lefcort et al. (1999) examined the competitive and predator avoidance behaviors of snails and tadpoles in outdoor mini-ecosystems with se diment from a metals-contaminated Superfund site (i.e., Pb, Zn, Cd). Previous investigations of aquatic invertebrates an d vertebrates yielded Pb tissue concentrations of 9 to 3800 mg/kg dw and 0.3 to 55 mg/kg dw, respectively. Several species interactions were studied in the pr esence of metal-contaminated sediment: Snails and tadpoles have similar dietary behavi ors. Thus, when placed in the same habitat they will compete for the same food items and ne gatively affect one another. However, when tadpoles exposed to a predator (i.e., through biweekly additions of 20 mL of water from tanks housing sunfish—10 mL from sunfish-fed snails, 10 mL from sunfish-fed tadpoles) were placed with snails, the tadpoles reduced sediment inge stion, while snails incr eased ingestion. Thus, snails were exposed to greater quant ities of metals in sediment. In an uncontaminated environment, snail r ecruitment (i.e., reproduction) was reduced in the presence of tadpoles. The addition of tadpol es increased the competition for food in the form of floating algae and the snails switched to feeding on algae that grew on the sediment. This decrease was due to competition alone. The eff ects on snail recruitment were even higher when tadpoles, the influence of a predator (i.e., sunfish extract), and metals in the sediment were all present. However, the predator effect was indi rect in that the tadpoles hid in the algal mats forcing the snails to feed primarily on the be nthic algae that grew on the sediment with highAX7-208 metal levels. Furthermore, although not significa nt, Pb levels in snails were higher when tadpoles and sunfish extract were present than when only metals in the sediment were present. Finally, snail predator avoidance was assesse d. Snails (control a nd lead-exposed) were stimulated with a predator indica tor (i.e., crushed snails and an ex tract of crushed snail). Control snails changed behaviors in the presence of the pr edator indicator, while exposed snails did not

alter their behavior. The author s suggested that metal exposure caused behavioral changes that alter competitive interactions and the perception of predators by the snails. Thus, Pb may affect the predator avoidance response of snails. In further study, Lefcort et al. (2000) examined the predator avoidance behaviors of snails and caddisflies. In separate experiments, the avoidance behavior of the snail, Physella

columbiana , and four caddisfly genera ( Agrypnia, Hydropsyche, Arctopsyche, Neothremma )

were evaluated. The snails were collected fr om reference lakes and lakes downstream of the Bunker Hill Superfund site. The snails from the affected lakes generally had higher cadmium, Pb, and zinc tissue levels implyi ng previous exposure to these me tals. Snail predator avoidance behavior was tested by exposure to crushed snail extract. Snails from the affected lakes did not reduce their activity when exposed to the snail extract, implying a reduced predator avoidance. The lack of response may make the snails at th e affected lakes more pr one to predation.

The caddisflies were evaluated at 36 sites from six different st reams. As with the snails,

the caddisflies from the affected streams had higher cadmium, Pb and zinc tissue levels. The time for caddisfly larvae to re spond (i.e., how long immobile) to disturbance (i.e., lifted from

water for 3 seconds and moved to a new locat ion) was evaluated. There was no correlation between tissue metal level and any response vari able (Lefcort et al., 2000). Therefore, the authors concluded that preexposure to metals did not reduce predator avoidance for caddisflies. Weber (1996) examined juvenile fathead minnows exposed to 0, 0.5, or 1.0 ppm Pb in water during a 2-week preexposure and 2-week testi ng period (4 weeks total exposure). Feeding behavior was evaluated by presenting tw o prey sizes (2-day-old and 7-day-old Daphnia magna ). Control fish began switching from larger, more difficult-to-capture 7-day-old daphnids to smaller, easier-to-catch 2-day-old prey by day 3. Lead-exposed fish displayed significant switching at day 3 (at 0.5 ppm) or day 10 (at 1.0 ppm). Thus, expos ure to Pb delayed the altering of prey size choices to less energetically costly prey.AX7-209 Lefcort et al. (1998) ex posed spotted frogs ( Rana luteiventris ) to 0.05 to 50 ppm Pb in water for 3 weeks. High levels of Pb reduced the fright response of tadpoles; suggesting a reduced avoidance of predators. Bullfrog larvae exposed to Pb in water (0.78 mg/L) and high or low dissolved oxygen were monitored for respiratory surfacing behavior (Rice et al., 1999). Larvae had a significantly increased number of trips to the water surface re gardless of oxygen content. Thus, the authors suggest that Pb may affect oxygen uptake such that larvae are under greater predation pressure due to increased time spent at the surface. Weis and Weis (1998) evaluated the e ffect of Pb expos ure on mummichog ( Fundulus

heteroclitus ) larvae prey capture rate, swimming behavior, and predator avoidance. Prey capture rates were affected after 4 weeks exposure at 1.0 mg Pb/L. The larvae were also more vulnerable to predation by grass shrimp ( Palaemonetes pugio ) at 1.0 mg Pb/L. Finally, the swimming behavior of mummichog larvae was affect ed at 0.3 and 1.0 mg Pb/L. Once the larvae were no longer exposed to Pb, they recovered their ability to capture prey and avoid predators. Clearly, exposure to Pb does affect the predator -prey interactions and th e ability of prey to avoid predators. The effect of Pb on these ecol ogical functions may alter community dynamics.

AX7.2.5.3 Summary The effects of Pb have primarily been studied in instances of point source pollution rather than area-wide atmospheric deposition; thus, the effects of atmos pheric Pb on ecological

condition remains to be defined. The evaluation of point source Pb within the EPA Ecological Condition Framework has been examined primar ily in relation to bi otic conditions. The

available literature focuses on st udies describing the effects of Pb in natural aquatic ecosystems with regard to community composition and species interactions. The effect s of Pb on the biotic condition of natural aquatic systems can be summarize d as follows: there is a paucity of data in the general literature that explores the effects of Pb in conjunction with all or several of the various components of ecological condition as de fined by the EPA. However, numerous studies are available associating the presence of Pb with effects on biotic conditions. In simulated microcosms or natural systems, environmental exposure to Pb in water and sediment has been shown to affect energy fl ow and nutrient cycling and benthic community structure. In field studies, Pb contamination has been shown to significantly alter the aquaticAX7-210 environment through bioaccumulation and alterati ons of community structure and function. Exposure to Pb in laboratory studies and simu lated ecosystems may alter species competitive behaviors, predator-prey interact ions, and contaminant avoidance be haviors. Alteration of these interactions may have negative effects on speci es abundance and community structure. In natural aquatic ecosystems, Pb is often found coexisting with other metals and other stressors. Thus, understanding the effects of Pb in natural systems is cha llenging given that observed effects may be due to cumulative toxicity from multiple stressors.AX7-211 

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