October 2006

EPA/600/R-05/144bF 

Air Quality Criteria for Lead 

Volume II of II 
EPA/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, NC
 II-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 reasonably may be expected to endanger public health or 

welfare; (2) to issue air quality criteria for them that assess the latest available scientific 

information on nature and effects of ambient exposure to them; (3) to set “primary” NAAQS to 

protect human health with adequate margin of safety and to set “secondary” NAAQS to protect 

against welfare effects (e.g., effects on vegetation, ecosystems, visibility, climate, manmade 

materials, etc); and (5) to periodically 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 information pertinent to Pb NAAQS development 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/m3 (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 since 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 Lead Staff Paper, prepared by EPA’s 

Office of Air Quality Planning and Standards (OPQPS), drew upon key findings and conclusions 

from 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 since the literature assessed in the 1986 Lead 

AQCD/Addendum and 1990 Supplement, with the main focus being on pertinent new 

information useful in evaluating health and environmental effects of ambient air lead exposures. 

This includes discussion in this document of information regarding: the nature, sources, 

distribution, measurement, and concentrations of lead in the environment; multimedia lead 

exposure (via air, food, water, etc.) and biokinetic modeling of contributions of such exposures 

to concentrations of lead in brain, kidney, and other tissues (e.g., blood and bone concentrations, 

as key indices of lead exposure).; characterization of lead health effects and associated exposureresponse relationships; and delineation of environmental (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 received were taken into account in making 

appropriate revisions and incorporating them into a Second External Review Draft (dated May, 

2006) which was released for further public comment 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 comments and CASAC advice received on these latter 

materials, as well as Second External Review Draft materials, were taken into account in making 

and incorporating further revisions into this final version of this Lead AQCD, which is being 

issued to meet an October 1, 2006 court-ordered deadline. Evaluations contained 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, ultimately, promulgation of decisions on potential 

retention or revision, as appropriate, of the current Pb NAAQS. 

Preparation of this document has been coordinated by staff of EPA’s National Center for 

Environmental Assessment in Research Triangle Park (NCEA-RTP). NCEA-RTP scientific 

staff, together with experts from academia, contributed to writing of document chapters. Earlier 
 II-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 valuable contributions provided by authors, contributors, and 

reviewers and the diligence of its staff and contractors in the preparation of this document. The 

constructive comments provided by public commenters 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 products 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, AND 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 (ENVIRONMENTAL EFFECTS OF LEAD).................................. AX7-1 
II-vii

Table of Contents 

Page

PREFACE......................................................................................................................................iii 

DISCLAIMER ................................................................................................................................ v 

List of Tables ................................................................................................................................. xi 

List of Figures..............................................................................................................................xiii 

Authors, Contributors, and Reviewers.......................................................................................... xv 

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 (ENVIRONMENTAL 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 Atmospherically Delivered Lead in 

Terrestrial Ecosystems....................................................... AX7-22 

AX7.1.2.1 Speciation of Atmospherically-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 of AtmosphericallyDelivered 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................................... AX7-61 
II-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 Natural 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 Bioavailability 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 Response/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.................................................... AX7-174 
II-ix

Table of Contents 

(cont’d) 

Page

 AX7.2.4 Exposure/Response of Aquatic Species........................... AX7-175 

AX7.2.4.1 Summary of Conclusions 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 Natural 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 .................................................................................................... AX7-211 
II-x

List of Tables 

Page

AX7-1.1.1 Relative Standard Deviation (RSD) for Lead Isotope Ratios on 

Selected Mass Spectrometers................................................................... AX7-2 

AX7-1.1.2 National Institute of Standards 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 Used 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 Used 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 Not 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 Current Acute Freshwater Criteria for Lead ............. AX7-109 

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

AX7-2.2.1 NAWQA Land Use Categories 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 Ambient 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 NAWQA Ambient Lead Levels in 

Whole Organism Tissues ..................................................................... AX7-138 
II-xi

List of Tables 

(cont’d) 

Page

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

AX7-2.3.2 Bioconcentration Factors for 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 Freshwater 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.................................................................................. AX7-200 
II-xii

List of Figures 

Number Page

AX7-1.1.1 Relationship of bioaccessibility (low, medium, high) versus 

speciation as shown with scanning electron micrographs of 

various Pb-bearing materials........................................................................ AX7-5 

AX7-1.1.2 Variation of bioavailability 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 species 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 ambient and natural levels of surface 

water dissolved lead (μg/L). .................................................................... AX7-131 

AX7-2.2.7 Spatial distribution of dissolved lead in surface water (N = 3445). ........ AX7-132 
II-xiii

List of Figures 

(cont’d) 

Number Page

AX7-2.2.8 Frequency distribution of ambient 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 ambient and natural levels of lead in 

liver tissue (μg/g dry weight)................................................................... AX7-137 

AX7-2.2.11 Frequency distribution of ambient 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 whole organism tissues (N = 332)............ AX7-140 

AX7-2.2.14 Lead cycle in an aquatic ecosystem......................................................... AX7-143 
II-xiv

Authors, Contributors, and Reviewers 

CHAPTER 4 ANNEX (TOXICOKINETICS, 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 27711 

 
II-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 Therapeutic 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 Corporation, 8191 Cedar Street, Akron, NY 14001 

Dr. N.D. Vaziri—Division of Nephrology and Hypertension, University of California – Irvine 

Medical Center, 101, The City Drive, Bldg 53, Room #125. Orange, CA 92868 

Dr. John Pierce Wise, Sr.—Maine Center for 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 Comparative and Environmental Toxicology, College of 

Veterinary Medicine, Cornell 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—National Center for Environmental Assessment (B243-01), 

U.S. Environmental Protection Agency, Research Triangle Park, NC 27711 

Dr. Michael Davis—National 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 12201 
II-xvi

Authors, Contributors, and Reviewers 

(cont’d) 

Contributors and Reviewers

(cont’d) 

Dr. Michael J. McCabe, Jr.—Dept of Environmental Medicine, University of Rochester, 

575 Elmwood Avenue, Rochester, NY 14642 

Dr. Theodore I. Lidsky—New York State Institute 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 Environmental 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 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 6 ANNEX (EPIDEMIOLOGICAL 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 27711 
II-xvii

Authors, Contributors, and Reviewers 

(cont’d) 

Principal Authors

Dr. David Bellinger—Children’s Hospital, Farley 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 Research 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, Room 1614, New York, NY 10032 

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

Dr. Stephen J. Rothenberg—CINVESTAV-IPN, Mérida, Yucatán, México & National Institute 

of Public Health, Cuernavaca, Morelos, Mexico 

Dr. Neal Simonsen—Louisiana State University Health Sciences Center, School of Public Health 

& Stanley S Scott Cancer Center, 1600 Canal 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, New York University School of 

Medicine, Tuxedo, NY 10987 
II-xviii

Authors, Contributors, and Reviewers 

(cont’d) 

Contributors and Reviewers

(cont’d) 

Dr. Kathryn Mahaffey—Office of Prevention, Pesticides and Toxic Substances, 

U.S. Environmental Protection Agency, Washington, DC 20460 

Dr. Karen Martin—Office of Air Quality Planning and Standards, U.S. Environmental 

Protection Agency, Research Triangle Park, NC 27711 

Dr. Zachary Pekar—Office of Air Quality Planning 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 (ENVIRONMENTAL 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., 1900 Minnesota Court, Suite 130, Mississauga, 

Ontario, L5N 3C9 Canada 

Dr. James Kaste—Department of Earth Sciences, Dartmouth College, 352 Main Street, 

Hanover, NH 03755 

Dr. John Drexler—Department of Geological Sciences, University of Colorado, 

1216 Gillespie Drive, Boulder, CO 80305 

Dr. Chris Johnson—Department of Civil and Environmental Engineering, Syracuse University, 

365 Link Hall, Syracuse, NY 13244 

Dr. William Stubblefield—Parametrix, Inc. 33972 Texas St. SW, Albany, OR 97321 
II-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 & Associates, 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 Environmental Health Sciences, 109 South Observatory, 

University of Michigan, Ann Arbor, MI 48109 

Dr. Judith Weis—Department of Biology, 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 Exposure Laboratory (D205-05), U.S. Environmental 

Protection Agency, Research Triangle Park, NC 27711 

Dr. Ginger Tennant—Office of Air Quality Planning and Standards, U.S. Environmental 

Protection Agency, Research Triangle Park, NC 27711 

Ms. Gail Lacey—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 
II-xx

U.S. Environmental Protection Agency Project Team 

for Development of Air Quality Criteria for Lead 

Executive Direction

Dr. Lester D. Grant (Director)—National Center 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)—National Center for Environmental Assessment 

(B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711 

Dr. James S. Brown—National Center for Environmental Assessment (B243-01), 

U.S. Environmental Protection Agency, Research Triangle Park, NC 27711 

Dr. Robert Elias—National Center for Environmental Assessment (B243-01), 

U.S. Environmental Protection Agency, Research 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—National 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 27711 
II-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 Specialist, National Center for Environmental 

Assessment (B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 

27711 (Retired) 

Ms. Emily R. Lee—Management Analyst, National Center for Environmental Assessment 

(B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711 

Ms. Diane H. Ray—Program Specialist, National Center for Environmental Assessment 

(B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711 

Ms. Donna Wicker—Administrative Officer, National Center for Environmental Assessment 

(B243-01), U.S. Environmental Protection Agency, Research Triangle Park, NC 27711 (Retired) 

Mr. Richard Wilson—Clerk, National Center for Environmental Assessment (B243-01), 

U.S. Environmental Protection Agency, Research Triangle Park, NC 27711 

Document Production Staff

Ms. Carolyn T. Perry—Task Order Manager, Computer Sciences Corporation, 2803 Slater Road, 

Suite 220, Morrisville, NC 27560 

Mr. John A. Bennett—Technical Information Specialist, Library Associates 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 Specialist, Library Associates of Maryland, 

11820 Parklawn Drive, Suite 400, Rockville, MD 20852 

Dr. Barbara Liljequist—Technical Editor, Computer Sciences Corporation, 2803 Slater Road, 

Suite 220, Morrisville, NC 27560 

Ms. Michelle Partridge-Doerr—Publications/Graphics Specialist, TEK Systems, 1201 Edwards 

Mill Road, Suite 201, Raleigh, NC 27607 

Mr. Carlton Witherspoon—Graphic Artist, Computer Sciences Corporation, 2803 Slater Road, 

Suite 220, Morrisville, NC 27560 
II-xxii

U.S. Environmental Protection Agency 

Science Advisory Board (SAB) Staff Office 

Clean Air Scientific Advisory Committee (CASAC) 

Chair

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

Albuquerque, NM 

Members

Dr. Joshua Cohen—Faculty, Center for the Evaluation of Value and Risk, Institute for Clinical 

Research and Health Policy Studies, Tufts 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, Department of Medicine, National Jewish Medical and 

Research Center, Denver, CO 

Dr. Bruce Fowler—Assistant Director for Science, Division of Toxicology and Environmental 

Medicine, Office of the Director, Agency for Toxic Substances and Disease Registry, U.S. 

Centers for Disease Control and Prevention (ATSDR/CDC), Chamblee, GA 

Dr. Andrew Friedland—Professor and Chair, Environmental Studies Program, Dartmouth 

College, Hanover, NH 

Dr. Robert Goyer [M.D.]—Emeritus Professor of Pathology, Faculty of Medicine, University of 

Western Ontario (Canada), Chapel Hill, NC 

Mr. Sean Hays—President, Summit Toxicology, Allenspark, CO 

Dr. Bruce Lanphear [M.D.]—Sloan Professor of Children’s Environmental Health, and the 

Director of the Cincinnati Children’s Environmental Health Center at Cincinnati Children’s 

Hospital Medical Center and the University of Cincinnati, Cincinnati, OH 

Dr. Samuel Luoma—Senior Research Hydrologist, U.S. Geological Survey (USGS), 

Menlo Park, CA 
II-xxiii

U.S. Environmental Protection Agency 

Science Advisory Board (SAB) Staff Office 

Clean Air Scientific Advisory Committee (CASAC) 

(cont’d) 

Members

(cont’d) 

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

Dr. Paul Mushak—Principal, PB Associates, 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 Division, 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 Health, Harvard University School of Public 

Health, Boston, MA 

Dr. Frank Speizer [M.D.]*—Edward Kass Professor of Medicine, Channing Laboratory, 

Harvard Medical School, Boston, MA 

Dr. Ian von Lindern—Senior Scientist, TerraGraphics Environmental Engineering, Inc., 

Moscow, ID 

Dr. Barbara Zielinska*—Research Professor, Division of Atmospheric Science, Desert Research 

Institute, Reno, NV 

SCIENCE ADVISORY BOARD STAFF 

Mr. Fred Butterfield—CASAC Designated Federal Officer, 1200 Pennsylvania Avenue, N.W., 

Washington, DC, 20460, Phone: 202-343-9994, Fax: 202-233-0643 (butterfield.fred@epa.gov) 

(Physical/Courier/FedEx Address: Fred A. Butterfield, 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 Administrator 
II-xxiv

Abbreviations and Acronyms 

αFGF 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 II 
II-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 α2 

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 auditory-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 index 
II-xxvi

BFU-E blood erythroid progenitor 

BLL blood lead level 

BLM biotic ligand model 

BM basement membrane 

BMI body mass index 

BDNF brain-derived neurotrophic 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 

45Ca calcium-45 radionuclide 

Ca-ATP calcium-dependent adenosine triphosphate 

Ca-ATPase calcium-dependent adenosine triphosphatase 

CaCO3 calcium carbonate 

CaEDTA calcium disodium ethylenediaminetetraacetic acid 

CAL calcitonin 

CaM calmodulin 

Ca-Mg-ATPase calcium-magnesium-dependent adenosine triphosphatase 

cAMP cyclic adenosinemonophosphate 

CaNa2 EDTA calcium disodium ethylenediaminetetraacetic acid 

CANTAB Cambridge Neuropsychological 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 B 
II-xxvii

CCD charge-coupled device 

CCE Coordination Center for Effects 

CCL carbon tetrachloride 

CCS cosmic calf serum 

C-CVRSA coefficient of component variance of respiratory sinus arrhythmia 

Cd cadmium 

109Cd cadmium-109 radionuclide 

CdU urinary cadmium 

CEC cation exchange capacity 

CESD, CES-D Center for Epidemiologic Studies Depression (scale) 

GFAP glial fibrillary acidic protein 

CFU-E colony forming unit blood-erythroid progenitor (cell count) 

CFU-GEMM colony forming unit blood-pluripotent progenitor (cell count) 

CFU-GM blood granulocyte/macrophage 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 protein 
II-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 

CVR-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 adenosine-3',5'-monophosphate 

DCV distribution of conduction velocities 

DEAE diethylaminoethyl (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 anodic stripping voltammetry 

dp/dt rate of left ventricular isovolumetric pressure 
II-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 

E2 estradiol 

EBE early biological effect 

EBV Epstein-Barr virus 

EC European Community 

EC50 effect concentration for 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-ethoxy-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 nitric oxide synthase 

EP erythrocyte protoporphyrin 

EPA U.S. Environmental Protection Agency 

Epi epinephrine 

EPMA electron probe microanalysis 

EPO erythropoietin 
II-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 

FEV1 forced expiratory volume in one second 

FGF fibroblast growth factor (e.g., βFGF, αFGF) 

FI fixed interval (operant conditioning) 

FIAM free ion activity model

FMLP N-formyl-L-methionyl-L-leucyl-L-phenylalanine 

fMRI functional magnetic resonance imaging 

FR fixed-ratio operant conditioning 

FSH follicle stimulating hormone 

FT3 free triiodothyronine 

FT4 free thyroxine 
II-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 microelectrodes 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 

gp91phox NAD(P)H oxidase 

GPT glutamic-pyruvic transaminase 

GPx glutathione peroxidase 
II-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 lipoprotein (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 

HNO3 nitric acid 

H2O2 hydrogen peroxide 

HOME Home Observation for Measurement of Environment 

HOS TE human osteosarcoma cells 
II-xxxiii

HPLC high-pressure liquid chromatography 

H3PO4 phosphoric acid 

HPRT hypoxanthine phosphoribosyltransferase (gene) 

HR heart rate 

HSI habitat suitability indices 

H2SO4 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 plasma atomic emission spectroscopy 

ICP-MS, ICPMS inductively coupled 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 deficit 
II-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 

LC50 lethal concentration at which 50% of exposed animals die 

LC74 lethal concentration at which 74% of exposed animals die 

LD50 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-NG-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 

LT50 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 covariance 
II-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 dinucleotide 
II-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-dependent 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-κB nuclear transcription factor-κB 

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 

NO2 nitrogen dioxide 

NO3 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 matter 
II-xxxvii

NORs nucleolar organizing regions 

NOS nitric oxide synthase; not otherwise specified 

NOx 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 

O2 oxygen 

ODVP offspring development 

OH hydroxyl 

7-OH-coumarin 7-hydroxy-coumarin 

1,25-OH-D, 1,25-OH D3 1,25-dihydroxyvitamin D 

24,25-OH-D3 24,25-dihydroxyvitamin D

25-OH-D3 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 

203Pb lead-203 radionuclide 

204Pb, 206Pb, 207Pb, 208Pb stable isotopes of lead-204, -206, -207, -208, respectively 
II-xxxviii

210Pb lead-210 radionuclide 

Pb(Ac)2 lead acetate 

PbB blood lead concentration 

PbCl2 lead chloride 

Pb(ClO4)2 lead chlorate 

PBG-S porphobilinogen synthase 

PBMC peripheral blood mononuclear cells 

Pb(NO3)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., PGE2, PGF2); 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 billion 
II-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 hormone-related protein 

PVC polyvinyl chloride 

PWM pokeweed mitogen 

PRWT progeny weight 

QA/QC quality assurance/quality control 

Q/V flux of air (Q) divided by volume of culture (V) 

r Pearson correlation coefficient 

R2 multiple correlation coefficient 

r

2

 correlation coefficient 

226Ra most stable isotope of radium 

R/ALAD ratio of ALAD activity before and after reactivation 

RAVLT Rey Auditory Verbal Learning Test 

86Rb 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 

222Rn most stable isotope of radon 

RNA ribonucleic acid 

ROS reactive oxygen species 

ROS 17.2.8 rat osteosarcoma cell line 
II-xl

RPMI 1640 Roswell Park Memorial Institute 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 

SO2 sulfur dioxide 

SOD superoxide dismutase 
II-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 Administration 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-thioguanine 
II-xlii

TGF transforming growth factor 

TH tyrosine hydroxylase 

232Th 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., TXB2) 

U uriniary 

235U, 238U 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-interval 
II-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 three components of matrix effects: 

atomic number (Z), absorption (A), and fluorescence (F) 

ZnNa2 DTPA zinc disodium diethylenetriaminepentaacetic acid 

ZnNa2 EDTA zinc disodium ethylenediaminetetraacetic acid 

ZPP zinc protoporphyrin 
 AX7-1 

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 further 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 perhaps 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 characterizing 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 contamination from an individual source(s) and its 

impact are of primary importance in risk assessment. The identification of exposure pathway(s) 

is fundamental to the risk analysis and critical in the planning of remediation scenarios. 

Although societies have been consuming Pb for nearly 9,000 years, production of Pb in 

the United States peaked in 1910 and 1972, at approximately 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 specific fingerprinting technique is required. It has been well 

established (Sturges and Barrie, 1987; Rabinowitz, 1995) that the stable isotope composition of 
 AX7-2 

Pb is ideally suited for this task. Lead isotopic ratio differences often allow multiple sources to 

be distinguished, with an apportionment of the bulk Pb concentration made to those sources. 

Lead has four stable isotopes: 204Pb, 206Pb, 207Pb, and 208Pb in natural abundances of 1.4, 

24.1, 22.1, and 52.4%, respectively. The radiogenic 206Pb, 207Pb, and 208Pb are produced by 

radioactive decay of 238U, 235U, and 232Th, respectively. Thus, the isotopic composition of Pb 

varies based on the U:Pb and Th:Pb ratios of the 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, generally in aqueous form, is analyzed on an 

ICP/MS (quadrapole, magnetic sector, or time-of-flight). Studies reviewing the most common 

analytical and sample preparation 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 Institute for Standards and Technology (NIST), 

Standard Reference Material (SRM) 981 and/or spiked 203Tl and 205Tl isotopes (Ketterer et al., 

1991; Begley and Sharp, 1997). The overall success 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 

204Pb:206Pb 0.0031 0.0032 0.00053 0.0011 

207Pb:206Pb 0.0032 0.0027 0.00053 0.00048 

208Pb:206Pb 0.0026 0.0024 0.00053 0.00046 

An extensive database comprising primarily North 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 apportion 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 studied: 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; Rabinowitz, 2005; Ketterer et al., 2001). 

AX7.1.1.2 Speciation in Assessing Lead Bioavailability in the Terrestrial Environment 

One of the three processes defined by the National Research Council in its review on 

bioavailability (NRC, 2002) is “contaminant interactions between phases”, more commonly 

referred to as “speciation.” 

A wide variety of analytical (XRD, EPMA, EXAFS, PIXE, XPS, XAS, SIMS) and 

chemical speciation modeling (SOILCHEM, MINTEQL, REDEQL2, ECOSAT, MINTEQA2, 

HYDRAQL, PHREEQE, WATEQ4F) tools have been used to characterize 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 speciation. Other techniques such as extended X-ray 

absorption fine structure (EXAFS) and extended X-ray absorption near edge spectroscopy 

(EXANES) show great promise and will be important in solving key mechanistic questions. 

In the case of phytotoxicity, the speciation of metals by direct measurement or chemical models 

of pore-water chemistry is most valuable. Further 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 physicochemical characteristics that are relevant to 

bioavailability. As a result of the direct impact these factors often have on a metal’s 

bioavailability, the term “bioaccessibility” has been adopted to define those factors. 

Speciation Role 

The accumulation of metals in the lithosphere 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 characteristic 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 bioavailability (the proportion of a 

toxin that passes a physiological membrane [the plasma 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 between phases (association-dissociation/bound-released), 

• transport of contaminants to organism, and 

• passage across a physiological membrane. 

As mentioned previously, the first process is more commonly referred to as speciation. 

The speciation of a toxic metal in the environment is a critical component of any ecosystem 

health risk assessment. Four important toxicologic and toxicokinetic determinants relating 

speciation to bioavailability are the (1) chemical form or species, (2) particle size of the metal 

form, (3) lability of the chemical 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 illustrated by in vivo and in vitro results for many of the 

common Pb-bearing minerals (Drexler, 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 these identified metal species (Davis et al., 1993; 

Ruby et al., 1992; Drexler and Mushak, 1995). 

Figure AX7-1.1.1. Relationship of bioaccessibility (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 solubility is not the only control for bioavailability, an increase in 

bioavailability has been directly attributed to a decrease in particle size: Barltrop and Meek 
 AX7-6 

(1979) observed that “the smaller the lead particle, 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 vitro results on numerous Pb-bearing phases ranging in particle size 

from 35 to 250 μm. While all phases studied showed increased bioavailability with decreasing 

particle size, more significantly, not all forms showed the same degree or magnitude of change 

(Figure AX7-1.1.2). Atmospheric particles are generally found to occur in bimodal populations: 

fine; 0.1 to 2.5 μm and coarse; 2.5 to 15 μm. This distribution is both 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 transport 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, Missouri. At this site, galena (PbS) was the 

dominant Pb species with mean particle-size distributions of 4, 6, and 14 μm in PM10 filters, 

house dust, and soils, respectively. These findings support the conclusion that aerial transport 

was the primary mechanism for Pb deposition in residential yards. Finally, such laboratory data 

have been supported by extensive epidemiologic evidence, enforcing the importance of particle 

size (Bornschein et al., 1987; Brunekreef et al., 1983; Angle et al., 1984). 

Particle Lability 

The impact on bioavailability of a metal particle’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 isolated from its environment. 

Data from several EPA Superfund sites and the Region VIII swine study (U.S. Environmental 

Protection Agency, 2004a) suggest that matrix associations, such as liberated versus enclosed, 

can play an important part in bioavailability. As illustrated in Figure AX7-1.1.3, two different 

media with similar total Pb concentrations and 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 sample 
 AX7-7 

Figure AX7-1.1.2. Variation of bioavailability with particle size. 

(BAF = 17%). Other evidence is more empirical, 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 can be useful in understanding the mechanistic 

phenomena controlling bioavailability. In addition, they will aid in developing and validating 

models to predict metal-environment interactions. 
 AX7-8 

Figure AX7-1.1.3. Illustration of particle lability and bioavailability 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 bioavailable 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, recontamination 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 bioavailability of a metal to plants from soils and sediments, it is 

generally assumed that both the kinetic rate of supply 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/sediment concentration is small, it is this phase 

that is directly available 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 for plants include (1) in-situ measurements 

using selective electrodes (Gundersen et al., 1992; 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 laboratory 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 Quantification and Speciation 

Bulk Quantification 

The major analytical methods most commonly used 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 stripping voltametry and differential pulse polarography). 

The choice of analytical method today for bulk quantification is generally ICP/AES or 

ICP/MS (U.S. Environmental Protection Agency, 2001). Since 1986, numerous SRMs have 

been developed for Pb (Table AX7-1.1.2), and several significant technological 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 (PM2.5) 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 detect 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 productivity increase), and increases precision. 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 either axial viewed plasma (horizontal), 

which provides greater sensitivity (DL= 0.8 μg Pb/L), or radial (vertical) viewed plasma, which 

performs best with high total dissolved 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, and Cr. The cells provide efficient interference 

(isobaric, polyatomic, and argide) removal independent of the analyte and sample matrix by 

using various reaction gases (H2, He, NH3), eliminating the need for interference correction 

equations. 

Speciation Tools 

A wide variety of analytical and chemical techniques have been 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; Szulczewski et al., 1997; Isaure et. al., 2002; Lumsdon and Evans, 

1995; Gupta and Chen, 1975; Ma and Uren, 1995; Charlatchka 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 looking for the proverbial “needle in a haystack.” 

Therefore, the speciation technique must not only provide the information outlined above, but it 

must also determine that information from a medium that contains very little of the metal. 

As illustrated in Figure AX7-1.1.5, for a Pb-contaminated 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 (physicochemical, extractive, and theoretical), 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 extractions, 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. These tools are briefly described below. 
 AX7-12 

Figure AX7-1.1.5. Bulk lead versus single species modality. Shaded area represents 

typical soil-lead concentrations found 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 tools 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 materials (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, availability, sample preparation, resolution), and usability as well as 

references to current applications. Although most of these tools are scientifically sound and 

offer important information on the mechanistic understanding of metal occurrence 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 selected 
 AX7-13 

material to describe the chemical or kinetic factors 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, monochromatic (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 information on both chemical state and atomic 

structure. Measurements are theoretically available for all elements and are not surface-sensitive 

nor sample-sensitive (i.e., gases, liquids, solids, and amorphous materials 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 particular oxidation state 

(Szulczewski et al., 1997; Shaffer et al., 2001; Dupont et al., 2002). The lower-energy spectra 

persist some 100 eV above the edge. These oscillations are termed EXAFS (extended X-ray 

absorption fine structure) and are more commonly 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 weighted average signal of structural configurations, providing 

information on the predominant form of the metal, while minor species, which may be more 

bioavailable, can be overlooked; (3) access to synchrotrons is limited and the beam time required 

to conduct a site investigation would be prohibitive; (4) a large 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 have 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 (nλ = 2dsinθ). The 

identification of a species from this pattern is based upon the position of the lines (in terms of 

θ or 2θ) and their intensities as recorded by an X-ray detector. The diffraction angle (2θ) is 

determined by the spacing between a particular set of atomic planes. Identification of the species 

is empirical, and current available databases 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 the 
 AX7-14 

diffraction lines and comparing them to standards, 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. Since this technique represents a bulk analysis, 

no particle size or lability information 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 heavy charged particles (generally He) to irradiate 

the sample. The resulting characteristic X-rays are emitted and detected in a similar manner as 

XRF, using Si-Li detectors. Particles generated from a small accelerator 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/cm2

); (2) no particle size information is provided; 

(3) peak overlaps associated with Si-Li detectors 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 quantification along with secondary electrons and/or backscatter electrons 

for visual inspection of a sample. Elements from beryllium to uranium can be nondestructively 

analyzed at the 50-ppm level with limited sample preparation. X-ray spectra can be rapidly 

acquired using either wavelength 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 inside the spectrometer housing so that Bragg’s law is satisfied 

and a particular wavelength can be focused. Photon energy pulses reflecting off the crystal are 

collected for an individual elemental line by the counter as a first approximation to 

concentration. For quantitative analysis, these intensities are compared to those of known 

standards and must be corrected for background, dead time, and elemental interactions (ZAF) 

(Goldstein et al., 1992). ZAF correction is in reference to the three components of matrix 

effects: atomic number (Z), absorption (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 heights approximately proportional 

to the quantum energy of the photon that generated them), producing a spectrum of energy 

(wavelength) versus counts. The net area under a particular peak (elemental line) is proportional 

to its concentration in the sample. For quantitative 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 relatively fast and inexpensive, available, 

and provides all of the required information for bioavailability assessments (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) are 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, time-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 beams, 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 identification 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 characterization 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 determining 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 excitation or photoionization of atoms within the near 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 discussed 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 extraction techniques that have been used extensively over the years 

(Tessier et al., 1979; Tessier and Campbell, 1988; Gupta and Chen, 1975). These methods use 

either a single or sequential extraction procedure to release species associated with a particular 

metal within the media. Single chemical extractions are generally used to determine the 

bioavailable amount of metal in a functional class: water-soluble, exchangeable, organically 

bound, Fe-Mn bound, or insoluble. 

In a similar approach, sequential extractions treat a sample with a succession of reagents 

intended to specifically dissolve different, less available 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 exchangeable, 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. These 
 AX7-17 Table AX7-1.1.3. Characteristics for Direct Speciation Techniques Tools 

Species Lability 

Species Particle 

Size 

Species Valance 

State 

Species Bonding 

Species 

Composition 

Species 

Abundance 

Element 

Specificity 

Isotopic 

Character 

Element 

Sensitivity 

Resolution 

XRD No No No No No# No No No 3-4 vol% Bulk EMPA Yes Yes Yes+ No Yes Yes? B-U No*** 50 ppm 0.5-1 μm SIMS No Yes No No Yes* Yes** Li-U Yes 1 ppb 10 μm XPS No No Yes Yes Yes* Yes** H-U No wt.% 100 μm XAS No No Yes Yes Yes* Yes** He-U No ppb 2 μm PIXIE No No No No Yes Yes** B-U No 10 ppm 4 μm * 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 and stoichiometric 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 phase. 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 correlated to bioavailability. 

Solution Speciation Using Computer-Based Models. Computer-based models are either 

based upon equilibrium constants or upon Gibb’s free energy values in determining metal 

speciation from solution chemistry conditions (concentration, pH, Eh, organic complexes, 

adsorption/desorption sites, and temperature). Both approaches are subject to mass balance and 

equilibrium conditions. These models have undergone a great deal of development in recent 

years, as reliable thermodynamic data has become available and can provide some predictive 

estimates of metal behavior. A good review of these models and their applications is provided 

by Lumsdon and Evans (1995). 

Speciation can be controlled by simple reactions; 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 strength, reaction kinetics are poorly 

known, and complex reactions with co-precipitation/adsorption are not modeled. 

The first limitation is perhaps the most significant for contaminated media. For example, 

none of the models would predict 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 equilibrium-based conceptual model that has been 

incorporated into regulatory agencies guidelines (including the EPA) to predict effects of metals 

primarily on aquatic biota and to aid in the understanding of their interactions 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 has not been applied to predict toxicity to soil 
 AX7-19 

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 principle, both be important. Second, it remains 

very difficult to univariately control the composition of the soil pore water and the metal 

concentrations in the pore water, due to re-equilibration of the system following modification of 

any of the soil properties (including addition of metal salts). 

Steenbergen et al. (2005) assessed acute copper toxicity to the earthworm Aporrectodea 

caliginosa using the BLM. To overcome the aforementioned problems inherent in soil toxicity 

tests they developed an artificial flow-through exposure system consisting of an inert quartz sand 

matrix and a nutrient solution, of which the composition 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 toxicity 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 methodologies to lower soil-Pb RBA (Brown et al., 

2003a,b). To date, the most common methods studied 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 fall within the categories of phosphate, biosolid, 

and Al/Fe/Mn-oxide amendments. 

Phosphate Amendments 

Phosphate amendments have been studied extensively 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; Yang et al., 2001; Ma et al., 1995). Research 

in this area stems from early work by Nriagu (1973) and Cotter-Howells and Caporn (1996), who 

pointed out the very low solubilities for many Pb-phosphates (Ksp !27 to !66), particularly 

chloropyromorphite [Pb5(PO4)3Cl]. The quest to transform soluble Pb mineralogical forms into 

chloropyromorthite continues to be the primary focus of most studies. Sources of phosphorous 
 AX7-20 

have included phosphoric acid (H3PO4), triple-super phosphate (TSP), phosphate rock, and/or 

hydroxyapatite (HA). Various studies 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 contain 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 dissolution rate of the original Pb species. 

A number of potentially significant problems associated with phosphate amendments have 

been recognized, including both phyto- and earthworm 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. (2004) 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 shown (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 into the groundwater) through competitive anion 

exchange. Some data (Lenoble et al., 2005) indicate that if one could amend arsenic and Pb 

contaminated soils with iron(III) phosphate this problem can be mitigated, however the increased 

concentrations of both phosphate and iron still exclude the application when drinking water is an 

issue. 

Biosolid Amendments 

Historically, biosolids have been used in the restoration of coal mines (Haering et al., 

2000; Sopper, 1993). More recently, workers have demonstrated the feasibility of their use in 

the restoration of mine tailings (Brown et al., 2003a), and urban soils (Brown et al., 2003b; 

Farfel et al., 2005). Mine tailings are inherently difficult to remediate 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; and (4) deficient in macro- and micronutrients. 
 AX7-21 

Stabilization (i.e., the establishment of a vegetative 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) demonstrated 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 healthy plant cover within 

2 years. Metal mobility was more difficult to evaluate. 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 (Agency for Toxic Substances and Disease Registry 

[ATSDR], 1988) and can be a significant 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 unless 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 commercial 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 remainder of the yard showed either 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 assortment 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 than 30 days. In vitro and in vivo data both 

indicated a 3 to 54% reduction in Pb RBA, with 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 transport is significantly accelerated in soils 

amended with biosolids (Al-Wabel et al., 2002; McBride et al., 1997, 1999; Lamy et al., 1993; 

Richards et al., 1998, 2000). Some of these studies indicate that metal concentrations in soil 

solutions up to 80 cm below the amended surface increased by 3- to 20-fold in concentration up 
 AX7-22 

to 15 years after biosolid application. The increase in metal transport is likely the result of 

elevated dissolved organic carbon (DOC) in the 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 Atmospherically Delivered Lead in 

Terrestrial Ecosystems 

The 1986 Lead AQCD (U.S. Environmental Protection Agency, 1986a) contains only a 

few minor sections that detail the speciation, distribution, and behavior of atmospherically 

delivered Pb in terrestrial ecosystems. The document concluded that the majority of Pb in the 

atmosphere at that time was from gasoline consumption: of the 34,881 tons of Pb emitted to the 

atmosphere in 1984, 89% was from gasoline use 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 consumption, but results largely from iron and steel 

foundaries, boilers and process heaters, the combustion of fossil fuels in automobiles, trucks, 

airplanes, and ships, and other 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 with ammonium halides (e.g., PbBrCl, PbBrCl2NH4Cl), 

while Pb emitted from smelters is dominated by Pb-sulfur species (Habibi, 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 surface of the earth (i.e., to vegetation and soils), 

are suspected to be in the form of PbSO4, 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 terrestrial systems, the different Pb sources, 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 hydrous Mn- and Fe-oxides (Miller and McFee, 

1983) for scavenging Pb in soils are discussed in some detail in the 1986 Lead AQCD. Nriagu 
 AX7-23 

(1974) used thermodynamics to argue that Pb-orthophosphates (e.g., pyromorphite) represented 

the most stable Pb phase in many soils and sediments. He further suggested that, because of the 

extremely low solubility of Pb-phosphate minerals, Pb deposition could potentially reduce 

phosphorous availability. Olson and Skogerboe (1975) reported that solid-phase PbSO4

dominated gasoline-derived Pb speciation in surface soils from Colorado, Missouri, and Chicago, 

while Santillan-Medrano and Jurinak (1975) suggested that Pb(OH)2, Pb(PO4)2, and PbCO3

could regulate Pb speciation in soils. However, insoluble organic material can bind strongly to 

Pb and prevent many inorganic phases from ever forming in soils (Zimdahl and Skogerboe, 

1977). 

The vertical distribution and mobility of atmospheric Pb in soils was poorly documented 

prior to 1986. Chapter 6 of the 1986 AQCD cited a few references suggesting that atmospheric 

Pb is retained in the upper 5 cm of soil (Reaves and Berrow, 1984). Techniques using radiogenic 

Pb isotopes had been developed to discern between gasoline-derived Pb and natural, geogenic 

(native) Pb, but these techniques were mostly applied to sediments (Shirahata et al., 1980) prior 

to the 1986 Lead AQCD. Without using these techniques, accurate determinations of the depthdistribution and potential migration velocities for atmospherically delivered Pb in soils were 

largely unavailable. 

Several technological advances, combined with the expansion of existing technologies 

after 1986 resulted in the publication of a large body of literature detailing the speciation, 

distribution, and geochemical behavior of gasoline-derived Pb in the terrestrial environment. 

Most notably, the development of selective chemical extraction (SCE) procedures as a rapid and 

inexpensive means for partitioning Pb into different soil and sediment phases (e.g., Pb-oxides, 

Pb-humate, etc.) has been exploited by a number of researchers (Tessier 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 coordination environment 

of Pb in soils, organic matter, organisms, and sediments (Manceau et al., 1996; Xia et al., 1997; 

Trivedi et al., 2003). X-ray absorption studies can be used for the in-situ determination 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 technology 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). Advances in voltammetric, diffusive gradients in 
 AX7-24 

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 species in the soils and solutions, researchers have 

used radiogenic Pb isotopes (206Pb, 207Pb, 208Pb) to quantify the distribution, speciation, and 

transport of anthropogenic Pb in soil profiles and in vegetation (Bindler et al., 1999; Erel et al., 

2001; Kaste et al., 2003; Klaminder 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 significant fraction of the “labile” 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 reservoirs (Friedland et al., 1992; Miller and 

Friedland, 1994; Johnson et al., 1995a; Wang and Benoit, 1997). Overall, a significant amount 

of research has been published on the distribution, speciation, 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 microorganisms remain elusive. 

AX7.1.2.1 Speciation of Atmospherically-delivered Lead in Terrestrial Ecosystems 

Lead in the Solid Phases 

Lead can enter terrestrial ecosystems through natural rock weathering 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, the occurrence of Pb as a minor 

constituent of rocks (ppm level), particularly granites, rhyolites, and argillaceous sedimentary 

rocks is the more pertinent source term for the vast majority of terrestrial ecosystems. During 

the hydrolysis and oxidation of Pb-containing minerals, divalent Pb is released to the soil 

solution where it is rapidly fixed by organic matter and secondary mineral phases (KabataPendias and Pendias, 1992). The geochemical form of natural Pb in terrestrial ecosystems will 

be strongly controlled by soil type (Emmanuel and Erel, 2002). In contrast, anthropogenically 

introduced Pb has a variety of different geochemical forms, depending on the specific source. 

While Pb in soils from battery reclamation areas can be in the form of PbSO4 or PbSiO3, Pb in 
 AX7-25 

soils from shooting ranges and paint spills is commonly found as PbO and a variety of Pb 

carbonates (Vantelon et al., 2005; Laperche et al., 1996; Manceau et al., 1996). Atmospherically 

delivered Pb resulting from fossil fuel combustion 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 anthropogenic Pb speciation in soils 

are not well understood, there are many studies that have partitioned Pb into its different 

geochemical phases. A thorough understanding of Pb speciation is critical in order to predict 

potential mobility and bioavailability (see Section AX7.1.1). 

Selective chemical extractions have been employed 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 using a weak acid or salt solution (e.g., BaCl2), followed 

immediately by an extraction targeting organic matter (e.g., H2O2 or NaOCl), further followed by 

an extraction targeting secondary iron oxides (e.g., NH2OH·HCl), and finally, a strong reagent 

cocktail (e.g., HNO3-HCl-HF) targets primary minerals. Tessier et al. (1979) developed this 

technique. More recently, this technique has been modified and developed specifically for 

different metals and different types 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, extractions are rarely specific to a single phase. For example, while 

H2O2 is often used to remove metals bound to organic matter in soils, others have demonstrated 

that this reagent destroys clay minerals and sulfides (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 documented to occur during sequential chemical 

extractions (Ho and Evans, 2000; Sulkowski and Hirner, 2006), and many reagents may not 
 AX7-26 

completely extract targeted phases. While chemical extractions provide some useful information 

on metal phases in soil or sediment, the results should be treated as “operationally defined,” e.g., 

“H2O2-liberated Pb” rather than “organic Pb.” 

Lead forms strong coordination complexes with oxygen on mineral surfaces and organic 

matter functional groups (Abd-Elfattah and Wada, 1981), because of its high electronegativity 

and hydrolysis constant. Therefore, 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 adsorbed to organic and inorganic soil particles strongly 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 removed 15% or less Pb in organic horizons from a montane forest 

in New Hampshire. The fact that relatively concentrated acids, reducing agents, oxidizing 

agents, or chelating agents are required to liberate the majority of Pb from soils is used as one 

line of evidence that Pb migration and uptake by plants 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 matter functional groups, or a phosphate solution, 

which chelates structural cations. Extractions used to destroy organic matter often rely on H2O2

or NaOCl. Both organic and mineral horizons typically have significant Pb in this soil phase. 

Miller and McFee (1983) used Na4P2O7 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 H2O2, 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 K4P2O7 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 montane forests in Vermont and New Hampshire. 
 AX7-27 

They found that repeated extractions with Na4P2O7 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 often used to quantify organically bound metals, this reagent can 

both disperse and dissolve Fe phases (Jeanroy and Guillet, 1981; Shuman, 1982). Acidified 

H2O2 has also been reported to destroy and release elements associated with secondary soil 

minerals (Papp et al., 1991; Ryan et al., 2002). 

Aside from organic forms, Pb is often found to be associated with secondary oxide 

minerals in soils. Pb can be partitioned with secondary oxides by a variety of mechanisms, 

including (1) simple ion exchange, (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 solutions of EDTA, oxalate, or hydroxylamine hydrochloride (HH). 

Miller and McFee (1983) used an EDTA solution followed by an HH solution to quantify Pb 

occluded by Fe and Mn minerals, respectively, in their 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 concluded that negligible Pb was associated with 

Mn-oxides and that 1 to 30% of the Pb was associated with Fe phases in the mineral soils in their 

study. Johnson and Petras (1998) reported that no Pb was removed 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 configuration of 

metals in untreated soil and sediment samples. This type of analysis has been extremely valuable 

for directly determining the coordination environment of Pb in a variety of soils and sediments. 

Since different elements have different electron binding energies (Eb), 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 edge 
 AX7-28 

represents Eb. The precise energy required to dislodge a core electron from a metal (i.e., Eb) will 

be a function of the oxidation state 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 energy region immediately after the absorption edge, X-ray absorption 

increases and decreases with a periodicity that represents the wave functions of the ejected 

electrons and the constructive and destructive interference with the wave functions of the nearby 

atoms. X-ray absorption studies used to investigate the periodicity of the absorption after Eb are 

referred to as EXAFS (extended X-ray absorption 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 speciation without some of the problems associated with chemical 

extractions (Bargar et al., 1997a,b, 1998). 

Manceau et al. (1996) used EXAFS to study soil contaminated by gasoline-derived Pb in 

France and found that the Pb was divalent and complexed to salicylate and catechol-type 

functional groups of humic substances. He concluded that the alkyl-tetravalent Pb compounds 

that were added to gasoline were 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 the relatively insoluble PbS when contaminated 

wetland soils were treated with sulfate. 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 hypochlorite) the added Pb appeared to adsorb instead to some form 

of SiO2. More recent XAS studies have demonstrated the importance of biomineralization 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 precipitation 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 environmental samples, numerous experimental-based 

XAS studies have documented in detail the coordination environment of Pb adsorbed to Feoxides, Mn-oxides, Al-oxides, and clay minerals (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 FeO6 octahedra on three different types of 

sites: on corners, edges, or faces. Ostergren 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 that the mechanism of Pb sorption to SiO2 surfaces is 

pH-dependent. In acid environments (pH < 4) Pb adsorbs largely as an inner-sphere mononulear 

complex, but as pH increases, Pb sorption increasingly occurs via the formation of surfaceattached covalent polynuclear Pb species. 

Recently, Jackson et al. (2005) used microfocused synchrotron-based X-ray fluorescence 

(ΦSXRF) to detail the distribution of Pb and Cu in the nematode Caenorhabditis elegans. They 

found that, while Cu was evenly distributed throughout the bodies of exposed C. elegans, Pb was 

concentrated in the anterior pharynx region. Microfocused X-ray diffraction indicated that the 

highly concentrated Pb region in the pharynx was actually comprised of the crystalline Pb 

mineral, pyromorphite. The authors concluded that C. elegans precipitated pyromorphite in the 

pharynx as a defense mechanism to prevent spreading the toxic metal to the rest of the 

organism’s body. They further suggested that, because 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 probably controlled by some 

combination of (a) Pb-mineral solubility equilibria, (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 dissolved Pb phases on soil organic matter. Dissolved 

Pb phases in soil solution can be some combination of Pb2+ and its hydrolysis species, Pb bound 

to dissolved organic matter, and Pb complexes with inorganic ligands such as Cl!

 and SO4

2!

. 

Alkaline soils typically have solutions supersaturated with respect to PbCO3, Pb3(CO3)2(OH)2, 

Pb(OH)2, Pb3(PO4)2, Pb5(PO4)3(OH), and Pb4O(PO4)2 (Badawy et al., 2002). Pb-phosphate 

minerals in particular are very insoluble, and calculations based 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 chelating agents, such as dissolved organic matter, 

can prevent the precipitation of Pb minerals (Lang and Kaupenjohann, 2003). 

Using a combination of desorption experiments 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), recrystallization-induced incorporation into the solid phase 

(Ainsworth et al., 1994), and formation and stabilization 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, suggesting Pb adsorption primarily to the substrates’ 

surfaces (Rouff et al., 2006). However, for Pb adsorption to calcite aging played a significant 

role due to detection of multiple sorption mechanisms, 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 and 8.2. At pH 9.4, no further uptake occurred with 

aging and very little desorption occurred. These results show the 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, respectively, 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 Pb2+ to the solid phase is also controlled by 

total metal loading, i.e., high Pb loadings will result in a lower fraction being partitioned to the 

solid phase. Sauvé et al. (1997, 1998) demonstrated that only a fraction of the total Pb in 

solution was actually Pb2+ in soils treated with leaf compost. The fraction of Pb2+ 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 Pb2+ in pore waters of agricultural and contaminated soils from Australia and the United 

States. In acidic soils, Al species can compete for sites on natural organic matter and inhibit Pb 

binding to surfaces (Gustafsson et al., 2003). 

Differential pulse anodic stripping voltammetry (DPASV) is a technique that is useful for 

identifying relatively low concentrations of Pb2+ and has found many applications in adsorption 
 AX7-31 

and partitioning experiments. This technique has been particularly useful for quantifying the Kd, 

or partitioning ratio of Pb in the solid-to-liquid phase (Kd = [total solid-phase metal in mg kgΓ1

] / 

[dissolved metal in mg L!1

]). While the exact Kd value is a function of pH, organic matter 

content, substrate type, total metal burden, and concentrations of competing ligands, such studies 

typically show that Pb has very strong solid-phase partitioning. Partitioning ratios determined by 

DPASV generally range from 103 to 106 in soils in the typical pH range (Sauvé et al., 2000b). 

Aualiitia and Pickering (1987) used thin film ASV to compare the 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-hydroxides, 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 Pb2+ but as bound to DOM. 

Sauvé et al. (2000a) compared the relative affinity of Pb2+ 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 Pb2+ that was available in solution 

than the leaf compost. Glover et al. (2002) used DPASV in studying the 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-delivered Lead in Terrestrial Ecosystems 

Radiogenic Pb isotopes offer a powerful tool for separating anthropogenic Pb from natural 

Pb derived from mineral weathering (Erel and Patterson, 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 radiogenic stable Pb isotopes (206Pb, 207Pb, and 208Pb) have a 

heterogeneous distribution in the earth’s crust primarily because of the differences in the halflives of their respective parents (238U, T1/2 = 4.7 × 109

 year; 235U, T1/2 = 0.7 × 109 year; 232Th, 
 AX7-32 

T1/2 = 14 × 109

 year). The result is that the ore bodies from which anthropogenic Pb are typically 

derived are usually enriched in 207Pb relative to 206Pb and 208Pb when compared with Pb found in 

granitic rocks. Graney et al. (1995) analyzed a dated core from Lake Erie, and found that the 

206Pb/207Pb value in sediment deposited in the 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 Emmanuel 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 mixture of both anthropogenic and geogenic Pb. 

Anthropogenic Pb has been detected relatively deep into 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 that total 20th-century industrial Pb input to their 

study site was approximately 0.4 g Pb mΓ2

. Lead concentrations in organic material were highest 

in the upper soil horizons, and decreased with depth. During snowmelt, Pb in the snowpack was 

mixed with the anthropogenic and geogenic Pb already in the topsoil, and spring melts contained 

a mixture of anthropogenic and geogenic particulate 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 210Pb also provides a tool for tracing atmospherically delivered Pb in 

soils. After 222Rn (T1/2 = 3.8 days) is produced from the decay of 226Ra (T1/2 = 1600 years), some 

fraction of the 222Rn escapes from rocks and soils to the atmosphere. It then decays relatively 

rapidly to 210Pb (T1/2 = 22.3 years), which has a tropospheric residence time of a few weeks 

(Koch et al., 1996). Fallout 210Pb 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 residence 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; Kaste et al., 2003). Atmospheric 210Pb (210Pbex hereafter,210Pb in 

“excess” of that supported by 222Rn in the soil) must be calculated by subtracting the amount of 
 AX7-33 

210Pb formed in soils by the in-situ decay of 222Rn from the total 210Pb (Moore and Poet, 1976; 

Nozaki et al., 1978). 

Benninger et al. (1975) measured fallout 210Pb in soils and streamwater at Hubbard Brook 

and at an undisturbed forest in Pennsylvania. They estimated atmospheric 210Pb export in 

streamwaters to be <0.02% of the standing 210Pb crop in the organic horizons. They used a 

simple steady-state model to calculate the residence time of Pb in the organic horizons to be 

5,000 years. This overestimate of the Pb residence 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 210Pb in the soil 

column in between. Dörr and Münnich (1989, 1991) used 210Pb profiles in soils of southern 

Germany to evaluate the behavior of atmospherically delivered Pb. They calculated the vertical 

velocity of Pb by dividing the relaxation depth (i.e., the depth at which 210Pb activity decreases to 

1/e, or approximately 37% of its surface value) by the 210Pb mean life of 32 years. They reported 

downward transit velocities of atmospherically deposited Pb at 0.89 ± 0.33 mm yearΓ1. The 

downward transport of atmospheric Pb was not affected by pH or soil type. However, since Pb 

velocities in the soil profile where identical to carbon velocities calculated with 14C, they 

concluded that Pb movement in forest soils is probably controlled by carbon transport. Kaste 

et al. (2003) used 210Pb to model the response time of atmospherically delivered Pb in the O 

horizon at Camel’s Hump Mountain in Vermont. They concluded that the forest floor response 

time was between 60 and 150 years, depending on vegetation zone and elevation. Using 

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

and into the mineral soil in the 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 

210Pb distributions to calculate the residence time of Pb in soils of Northern Sweeden. They 

concluded that atmospheric Pb had a residence time of ~200 years in the mor layer of mature 

forests. They also provided evidence suggesting that forests in an earlier stage of ecological 

succession have a much shorter Pb residence 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 (forest 
 AX7-34 

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) demonstrated 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 that Pb concentrations 

in the litter layer (fresh litter + Oi horizon) sampled in a transect from Vermont to Quebec 

decreased significantly between 1979 and 1996, reflecting 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 would reach steady state with respect to Pb 

(1.3 μg gΓ1 Pb) by 2100. Both suggested that the vertical movement of organic particles 

dominated Pb transport in the soil profile. 

AX7.1.2.3 Inputs/Outputs of Atmospherically-delivered Lead in Terrestrial Ecosystems 

The concentration of Pb in contemporary rainfall 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 elemental concentrations (including 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 decrease 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 averaged 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 estimated 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, wet, 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 (Yi 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 dry 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 relative 
 AX7-36 

contributions of dry deposition may have changed in the last few decades. If the major source of 

Pb to a terrestrial ecosystem is resuspended particulates from transportation corridors, then the 

particle size fraction that dominates deposition may be relatively coarse (>50 μm) relative to 

other atmospheric sources (Pirrone et al., 1995; Sansalone et al., 2003). 

Total contemporary loadings to terrestrial ecosystems are approximately 1 to 2 mg mΓ2

yearΓ1 (Wu et al., 1994; Wang et al., 1995; Simonetti et al., 2000; Sabin et al., 2005). This is a 

relatively small annual flux of Pb if compared to the reservoir of approximately 0.5 to 4 g mΓ2 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 the base level of clouds often have higher burdens of atmospheric 

contaminants (Siccama, 1974). A Pb deposition model by Miller and Friedland (1994) predicted 

2.2 and 3.5 g Pb mΓ2

 deposition for the 20th century in the deciduous zone (600 m) and the 

coniferous zone (1000 m), respectively. More recently, Kaste et al. (2003) used radiogenic 

isotope measurements on the same mountain to confirm higher loadings at higher elevation. 

They measured 1.3 and 3.4 g gasoline-derived Pb mΓ2 in the deciduous zone and coniferous 

zones, respectively. Higher atmospheric Pb loadings to higher elevations are attributed to 

(1) the higher leaf area of coniferous species, which are generally more prevalent at high 

elevation; (2) higher rainfall at higher elevation; and (3) increased cloudwater impaction at high 

elevation (Miller et al., 1993). 

Although inputs of Pb to ecosystems are currently 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 from more biologically 

active parts of the watershed (unless later disturbed or redistributed, e.g., by flooding, dredging, 

etc.). Seeps and streams at the HBEF have Pb concentrations on the order of 10 to 30 pg Pb gΓ1

(Wang et al., 1995). At a remote valley in the Sierra Nevada, Pb concentrations in streamwaters 

were on the order of 15 pg Pb gΓ1 (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). Tyler 
 AX7-37 

(1981) noted that Pb losses from a spruce forest 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 movement 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; Marsh and Siccama, 1997; Yanai et al., 2004; Murray et al., 

2004). While some of this contaminant Pb is attributed to paint, salvage yards, shooting ranges, 

and the use of Pb arsenate as a pesticide in localized areas (Francek, 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 incineration (Newhook et al., 

2003; Polissar et al., 2001). Surface soils in Michigan, 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 upper soil horizons are 20 to 

40 ppm Pb, and approximately 75% of this Pb has 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 during the 20th century. Total Pb deposition during 

the 20th century has been estimated at 1 to 3 g Pb mΓ2

, 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 mΓ2

 (Friedland et al., 1992; 

Miller and Friedland, 1994; Johnson et al., 1995a; Kaste et al., 2003; Yanai et al., 2004; Evans 

et al., 2005). The pool of Pb in above- and below-ground biomass at the HBEF is on the order of 

0.13 g Pb mΓ2 (Johnson et al., 1995a). 

The amount of Pb that has leached into mineral 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 mΓ2

 (out of 1 to 2 g Pb mΓ2 in the total soil profile), at higher elevations the thicker coniferous 

forest floor retained more than 90% of the total 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 mΓ2

. 
 AX7-38 

Compared to their estimated 20th-century atmospheric Pb deposition of 0.9 g Pb mΓ2

, 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 significant component of the labile Pb at the HBEF. 

They calculated that Pb fluxes to the HBEF by atmospheric 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 underneath formerly plowed land in Rhode Island, 

New Hampshire and Connecticut to assess the depth-distribution of atmospheric Pb. They 

reported that 65% of the atmospheric Pb deposited 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 present in ecosystems in a variety of different 

biogeochemical phases. A combination of Pb adsorbtion processes and the precipitation of Pb 

minerals will typically keep dissolved Pb species 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 contaminant Pb (Nriagu, 1974; Ruby et al., 1994; 

Jackson et al., 2005), the influence of organic matter 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 phases (Manceau et al., 1996; Xia et al., 1997; Lang 

and Kaupenjohann, 2003). As the abundance of organic matter declines in soil, Pb adsorption to 

inorganic soil minerals and the direct precipitation of Pb phases 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, distribution, uptake, and fluxes of atmospherically delivered Pb in 

ecosystems. Among other things, these studies have shown that industrial Pb represents a 

significant fraction of total labile Pb in watersheds. 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 of 
 AX7-39 

these studies have provided compelling evidence that the biomineralization of Pb phosphates by 

soil organisms can play an important role in the 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 systems during the 20th century. The migration and 

biological uptake of Pb in ecosystems is relatively low. The different biogeochemical behaviors 

of Pb reported by various studies may be a result of the many different analytical techniques 

employed, or they may be a result of natural variability in the behavior of Pb in different 

systems. 

Lead Uptake into Plants 

Plants take up Pb via their foliage and through 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. Environmental 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; Angelova et al., 2004). Data examined from 

experimental grassland plots in southeast England demonstrated that atmospheric Pb is a greater 

contributor than soil-derived Pb in crop plants 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 atmosphere, although uptake from soil was also 

observed. The concentration of Pb in tobacco seeds was linearly related to the concentration of 

Pb in the exchangeable and carbonate-bound fractions of soil, as measured using sequential 

extraction (Angelova et al., 2004). Lead in soil is more significant when considering uptake into 

root vegetables (e.g., carrot, potato), since, as was noted in the 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 through the cell membranes of root hairs; this is the 

mechanism of uptake for water and nutrients. The apoplastic route is an extracellular route 

between epidermal cells into the intercellular spaces of the root cortex. Previously, Pb was 
 AX7-40 

thought to enter the plant via the symplastic route, probably by transport 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 was 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 the 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 the binding of Pb ions at root surfaces and cell 

walls. In a study by Wierzbicka (1999), 21 different plant species were exposed to Pb2+ in the 

form of Pb-chloride. The plant 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 roots. Only a small amount of Pb was translocated (~2 to 4%) to the shoots 

of the plants. Lead in roots was present in the deeper layers of root tissues (in particular, the root 

cortex) and not only on the root surface. There was no correlation between Pb tolerance 

(measured as root mass increase expressed as a percentage 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 the onion, were more tolerant to Pb than plants 

developing from seeds. This tolerance was assumed 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 modeled as part of Eco-SSL development (U.S. 

Environmental Protection Agency, 2005a). The relationship derived between Pb in the soil and 

Pb in a plant was taken from Bechtel Jacobs 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 recognizes that the ratio of Pb concentration in plant to Pb 

concentration in soil is not constant. 

Invertebrates 

There was no clear evidence suggesting a differential uptake of Pb into different species 

of earthworm (Lumbricus terrestris, Aporrectodea rosea, and A. caliginosa) collected around a 

smelter site near Avonmouth, England (Spurgeon and Hopkin, 1996a). This 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 could be due to different 

feeding behaviors, as Lumbricus feeds on organic material and Apporectodea species are 

geophagus, ingesting large amounts of soil during feeding. The differences between species also 

may be related to differing efficiencies in excretory mechanisms (Pižl and Josens, 1995). 

However, the interpretation of species difference is complicated by a number of potentially 

confounding variables, such as soil characteristics (e.g., calcium or other nutrient levels) 

(Pižl and Josens, 1995). 

The bioaccumulation of Pb from contaminated 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 (Davies et al., 2003). This coincided with the 

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

concentration of Pb in worms to the concentration 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 contamination (<3000 mg/kg) 

until a critical concentration is reached, at which point this mechanism breaks down, resulting in 

unregulated accumulation and mortality. This study was conducted using test methods where 
 AX7-42 

(AX7-2) 

soil was not allowed to equilibrate following the addition 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 conditions (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-nitrate was more toxic 

than Pb-chloride to survival and reproduction of the springtail Folsomia candida. However, 

percolation (removal of the chloride or nitrate counterion) caused a significant decrease in Pbnitrate toxicity such that there was no difference in toxicity once the 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 nitrate 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 Agency, 2005a). The relationship derived between 

Pb in the soil and Pb in an earthworm was taken 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 weight) and Csoil is the 

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

worm to Pb concentration in soil is not constant. 

Wildlife 

Research has been conducted to determine what Pb concentrations in various organs 

would be indicative of various levels 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), Columbiformes (e.g., doves, pigeons), and Galliformes 

(e.g., turkey, pheasant, partridge, quail, chickens). The three levels of effect 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) toxic, a threshold level marking the initiation of clinical 

signs, such as anemia, lesions in tissues, weight loss, muscular incoordination, green diarrhea, 

and anorexia; and (3) compatible with death, an approximate threshold value associated with 
 AX7-43 

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 subclinical levels in Table AX7-1.2.1 should be 

considered “background” (Franson, 1996). Levels in the subclinical range are indicative of 

potential injury from which the bird would probably recover if Pb exposure was terminated. 

Toxic residues could lead to death. Residues above the compatible-with-death threshold are 

consistent with Pb-poisoning mortality (Franson, 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 Franson (1996). Scheuhammer (1989) found blood Pb 

concentrations of between 0.18 and 0.65 μg/mL 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 been correlated with toxicity 

(Ma, 1996). The tissues commonly analysed for 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 levels of 3 to 19 mg/kg dw. Ma (1996) concluded that Pb levels 

less than 5 mg/kg dw in liver and 10 mg/kg dw in kidney were not associated with toxicity, but 

that levels greater than 5 mg/kg dw in liver and greater than 15 mg/kg dw in kidney could be 

taken as a chemical biomarker of toxic exposure to Pb in mammals. Humphreys (1991) noted 

that the concentrations of Pb in liver and kidney 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 Agency, 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 for plants (Eq. 8-1) and 

earthworms (Eq 8-2), the equation for mammalian 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 tolerance in plants and invertebrates have been 

described, including avoidance (i.e., root redistribution, food rejection), exclusion (i.e., selective 

uptake and translocation), immobilization at the 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 detoxification by regulating internal Pb concentrations 

through the formation of Pb-containing 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 content in the root and root exudates was increased in Pb-tolerant varieties. 

The authors suggested that the oxalate reduced Pb bioavailability, and that this was an important 

tolerance mechanism (Yang et al., 2000). Sharma et al. (2004) found Pb-sulfur 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 solution. They hypothesized that these sulfur 

ligands were indicative of glutathione and phytochelatins, which play a role in heavy metal 

homeostasis and detoxification (Sharma et al., 2004). 

Sea pinks (Armeria maritima) grown on a metal-contaminated site (calamine spoils more 

than 100 years old) accumulated 6Η 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 greater accumulation of Pb into older leaves was not 

observed in plants grown hydroponically in the laboratory. The authors hypothesized that this 

sequestering of Pb into the oldest leaves was a detoxification mechanism (Szarek-Lukaszewska 

et al., 2004). 

Terrestrial invertebrates also mitigate Pb toxicity. 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 enzymes (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 that 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). Gintenreiter et al. (1993) found that Lepidoptera 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 (Pb5(PO4)3Cl), was localized in the anterior pharynx 

region of the nematode Ceanorhabditis elegans (Jackson et al., 2005). The authors hypothesized 
 AX7-46 

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 that these organisms may 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 decreased blood ALAD activity in birds and mammals 

exposed to Pb (U.S. Environmental Protection 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, and production of fatty 

acids. Effects on growth are covered in Section AX7.1.4. 

Biochemically, Pb adversely affects hemoglobin synthesis in birds and mammals. Early 

indicators of Pb exposure in birds and mammals include decreased blood ALAD concentrations 

and increased protoporphyrin IX activity. The effects 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 ecologically 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 inhibition 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 sublethal 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 protoporphyrin levels in the blood. As found in 

other studies, ALAD activity was the most sensitive indicator of Pb exposure, decreasing to 3% 
 AX7-47 

of the reference value at a blood Pb concentration 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 related to Pb exposure, 

although inhibition of blood ALAD and increased protoporphyrin concentrations were measured 

in ospreys. Henny et al. (1991) hypothesized that 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 effects of Coeur d’Alene River sediment on 

waterfowl, focusing on mallard ducklings for 6 weeks after hatching. The study revealed that a 

90% reduction in ALAD activity and a greater 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 found 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 concentrations in goslings were 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 activity and a 4-fold increase in protoporphyrin 

concentration, similar to conditions found in the ducklings, although blood Pb concentrations 

were half those found in the ducklings. More serious effects were seen in the goslings when 

blood Pb reached 2.52 mg/kg, including decreased growth and 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 mortality nor clinical signs of toxicity. A dose of 6.55 mg/kg-day 

resulted in blood Pb levels of 1.58 μg/mL. There 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 in 
 AX7-48 

plasma levels of uric acid and creatinine 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 exposure. 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 lipid peroxides and changes in glutathione 

concentrations, which may be related to peroxidative damage of cell membranes (Mateo and 

Hoffman, 2001). There are species-specific differences in resistance to oxidative stress (lipid 

peroxidation), which may explain why Canada geese are more sensitive to Pb poisoning than 

mallards (Mateo and Hoffman, 2001). Lead also caused 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 enough that birds decreased their food intake. 

Lead also induces lipid peroxidation in plants. Rice plants exposed to a highly toxic level 

of Pb (1000 μM in nutrient solution) showed elevated levels of lipid peroxides, increased activity 

of superoxide dismutase, guaiacol peroxidase, ascorbate peroxidase, and glutatione reductase 

(Verma and Dubey, 2003). The elevated levels of these enzymes suggest the plants may have an 

antioxidative defense mechanism against oxidative injury caused by Pb (Verma and Dubey, 

2003). 

AX7.1.2.6 Factors that Modify Organism Response 

Research has demonstrated that Pb may 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 factors that modify responses of organisms to Pb 

are described in the following sections. 

Genetics 

Uptake and toxicity of Pb to plants are influenced strongly by the type of plant. Liu et al. 

(2003) found that Pb uptake and translocation by rice plants differed by cultivar (a cultivated 
 AX7-49 

variety of plant produced by selective 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 concentrations in the grains at the ripening stage. 

This study also found that toxicity varied by cultivar; 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 dams. Lead delayed the timing of puberty and 

suppressed hormone levels in F344 offspring. These effects were not observed in the offspring 

of SD rats, even when the dose was doubled. The authors conclude that F344 rats are more 

sensitive to Pb (Dearth et al., 2004). 

Biological Factors 

Several biological factors may influence Pb uptake and organism response, including 

organism age, sex, species, feeding guild, and, for 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 birth, such as songbirds), are considered more 

sensitive than precocial birds (species that have a high degree of independence at birth, such as 

quail, ducks, and poultry) (Scheuhammer, 1991). 

Gender can also have an effect on the accumulation of Pb by wildlife (Eisler, 1988). 

Female birds accumulate more Pb than males (Scheuhammer, 1987; Tejedor and Gonzalez, 

1992). These and other authors have related this to the increased requirement for calcium in 

laying females. 

Different types of invertebrates accumulate different 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 invertebrates, 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-building spider, and this may be due to the more 
 AX7-50 

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 detoxify and excrete excess metals more effectively 

than males (Wilczek et al., 2004). Lead accumulation has been measured in numerous species of 

arthropods with different feeding 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 season (Rabitsch, 1995a). 

Uptake of Pb may be enhanced by symbiotic associations between plant roots and 

mycorrhizal fungi. Similar to the mechanism associated 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 particular ions, and the transfer of ions through the 

soil (U.S. Environmental Protection Agency, 1986a). There have been contradictory results 

published in the literature regarding 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 the availability varied 

with Pb concentration in soil. Bioavailability was measured as the soluble plus exchangeable Pb 

fraction from sequential extraction analysis. The authors hypothesized 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 stimulates microbial growth (Lin et al., 2004). 

Increased bioavailability of Pb in soil may increase the uptake of Pb into plants, although 

this was not assessed by Lin et al. (2004). However, Dixon (1988) found that red oak 

(Quercus rubra) seedlings with abundant ectomycorrhizae had lower Pb concentrations in their 

roots than those seedlings without this fungus, although only at the 100 mg Pb/kg sandy loam 

soil concentration (no differences were found at lower 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 toxicity 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 determinant of Pb body burdens in small mammals, with 

insectivorous animals accumulating more Pb than herbivores or granivores (U.S. Environmental 
 AX7-51 

Protection Agency, 1986a). In fact, the main issue identified by the EPA (U.S. Environmental 

Protection Agency, 1986a) related to invertebrate uptake of Pb was not toxicity to the 

invertebrates, but accumulation of Pb to levels that may be toxic to their consumers. Several 

authors suggest that shrews are a good indicator of metal contamination, because they tend to 

accumulate higher levels of metals than herbivorous 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 soil (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 chemical form of Pb, the presence of other metal 

ions, soil type, soil pH, cation exchange capacity (CEC), the amount of Fe/Mn-oxide films 

present, organic matter content, temperature, light, 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 influence on its toxicity to plants. For 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 concentration range of 0 to 5000 mg/kg, root growth was 

inhibited by 24% at 500 mg/kg for Pb-chloride and an EC50 of 2400 mg/kg was calculated from a 

dose-response curve. Plant growth ceased at 5000 mg/kg and shoots exhibited an EC50 of 

2800 mg/kg. For Pb-oxide exposure (concentration range of 0 to 10,000 mg/kg), reported results 

indicate an EC50 of 12,000 mg/kg for shoot growth and an EC50 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 with 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 elevated 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 concentrations after 3 months of growth 

compared to plants grown at pH 6.4 (Allinson and Dzialo, 1981). 

Invertebrates 

The uptake of Pb into invertebrates depends on the physical environment and parameters 

such as pH, calcium concentration, organic matter 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 soil 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 relationship 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 factors (ratio of Pb in 

worm to Pb in soil) were lower at higher Pb concentrations in soil. The authors attribute some of 

this to a lowered feeding activity in worms at higher Pb concentrations (Bengtsson et al., 1986). 

Beyer et al. (1987) and Morgan and Morgan (1988) 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 evaluated worm uptake of Pb relative 

to pH, soil calcium concentration, and organic matter content. Morgan and Morgan (1988) also 

considered CEC, and Beyer et al. (1987) considered concentrations of phosphorus, potassium, or 

magnesium in soil. Both studies found that calcium concentrations in soil were correlated with 

soil pH. Morgan and Morgan (1988) also found that CEC was correlated with percentage 

organic matter. Soil pH (coupled with CEC) and 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 concentrations of phosphorus in soil had no effect. 
 AX7-53 

Nutritional Factors 

Diet is a significant modifier of Pb absorption and of toxic effects in many species of 

birds and mammals (Eisler, 1988). Dietary deficiencies in calcium, zinc, iron, vitamin E, copper, 

thiamin, phosphorus, magnesium, fat, protein, minerals, and ascorbic acid increased Pb 

absorption and its toxic effects (Eisler, 1988). 

Mateo et al. (2003b) studied intraspecies 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; however, it did not 

alleviate any sign of the Pb poisoning. The authors hypothesize that inhibition of antioxidant 

enzymes and interaction with sulfhydryl groups of proteins may have a greater influence on Pb 

toxicity than lipid peroxidation (Mateo et al., 2003b). The effects observed by Mateo et al. 

(2003b) were associated with very high concentrations of Pb in diet (1840 mg Pb/kg diet), much 

higher than would be found generally in the environment, and high enough that the birds 

decreased their food intake. 

Mallard ducklings were exposed to Pb-contaminated sediment and either a low nutrition 

or optimal nutrition diet (Douglas-Stroebel et al., 2005). Lead 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 diet alone. These effects may be due to the lownutrition diet being deficient in levels of protein, amino acids, calcium, 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 either a low- or high-calcium diet (Snoeijs et al., 2005). Lead 

uptake into tissues was enhanced by a low-calcium 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 increased susceptibility of females to Pb (Snoeijs 

et al., 2005). 

Interactions with Other Pollutants 

Lead can interact with other pollutants to exert toxicity in an antagonistic (less than 

additive), independent, additive, or synergistic (more than additive) manner. Concurrent 

exposure to Pb and additional pollutant(s) can affect the ability of plants to uptake Pb or the 

other pollutant. However, the uptake and toxic response of plants exposed to Pb combined with 
 AX7-54 

other metals is inconsistent (Påhlsson, 1989). Therefore, no generalizations 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 + copper, 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 not consistent with the response 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, antagonism 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 (present 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 determine whether they could change the 

inhibition of ALAD in liver, kidney, 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 hypothesized that cadmium may block or antagonize 

the intestinal absorption of Pb, or the metallothionein induced by cadmium may sequester Pb. 

However, this was not observed in pigs, where 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 (Phillips et al., 2003). 
 AX7-55 

AX7.1.2.7 Summary 

The current document expands upon and updates knowledge related to the uptake, 

detoxification, physiological effects, 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 conclusions 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 documented previously that Pb in soil also is taken 

up by plants, although most remains in the roots, there is little translocation to shoots, leaves, or 

other plant parts (U.S. Environmental Protection Agency, 1986a). More recent work continues 

to support this finding (Sieghardt, 1990), and one study found increased 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 the symplastic route (through 

cell membranes), although it was unknown whether some Pb also may be taken up via the 

apoplastic route (between cells) (U.S. Environmental Protection Agency, 1986a). Recent work 

has shown that the apoplastic route of transport is stopped in the primary roots by the endodermis 

(Sieghardt, 1990), supporting the previous conclusion that the symplastic route is the most 

significant route of transport into plant cells. 

Species Differences in Uptake into Earthworms 

Different species of earthworm accumulated different amounts of Pb, and this was not 

related to feeding strategy (U.S. Environmental Protection Agency, 1986a). This is supported by 

recent work, which has shown Aporrectodea accumulated more than Lumbricus (Terhivuo et al., 

1994; Pižl and Josens, 1995), although this is not consistently 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 uptake and toxicity to plants and invertebrates (U.S. Environmental 

Protection Agency, 1986a). The oxide form is less toxic that the chloride or acetate forms, 

which are less toxic than the nitrate form of Pb (Khan and Frankland, 1983; Lock and Janssen, 

2002; Bongers et al., 2004). However, these results must be interpreted with caution, as the 

counterion (e.g., the nitrate ion) may be contributing to the observed toxicity (Bongers et al., 

2004). 

Detoxification in Plants 

Lead may be deposited in root cell walls as a detoxification 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 soil, and that this was an important tolerance 

mechanism. Other hypotheses put forward recently include the presence of sulfur ligands 

(Sharma et al., 2004) and the sequestration of Pb in old leaves (Szarek-Lukaszewska et al., 2004) 

as detoxification mechanisms. 

Detoxification in Invertebrates 

Lead detoxification has not been studied extensively in invertebrates. Glutathione 

detoxification enzymes were measured in two species of spider (Wilczek et al., 2004). Lead may 

be stored in waste nodules in earthworms (Hopkin, 1989) or as pyromorphite in the nematode 

(Jackson et al., 2005). 

Physiological Effects 

The effects on heme synthesis (as measured by ALAD activity and protoporphyrin 

concentration, primarily) have been well-documented (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 damage 
 AX7-57 

other organs, have been reported (Brar et al., 1997a,b). Lead also may cause lipid peroxidation 

(Mateo and Hoffman, 2001), which may be alleviated 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, physical/environmental factors, nutritional 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 (Dearth et al., 2004). Younger animals are more 

sensitive than older animals (Eisler, 1988; Scheuhammer, 1991), and females generally are more 

sensitive than males (Scheuhammer, 1987; Tejedor 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 herbivores (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 deficient in nutrients (including calcium) result in increased uptake 

of Pb (Snoeijs et al., 2005) and greater toxicity (Douglas-Stroebel et al., 2005) in birds, relative 

to diets containing adequate nutrient levels. 

Mycorrhizal fungi may ameliorate Pb toxicity until a threshold is surpassed (Malcova and 

Gryndler, 2003), which may explain why some studies show increased uptake into plants (Lin 

et al., 2004) while others show no difference or less uptake (Dixon, 1988). Uptake of Pb into 

plants and soil invertebrates increases with a decrease in soil pH. However, calcium content, 

organic matter content, and cation exchange capacity of soils also have had a significant 

influence on uptake of Pb into plants and invertebrates (Beyer et al., 1987; Morgan and Morgan, 

1988). 

Interactions of Pb with other metals are 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 that modify terrestrial 
 AX7-58 

organism responses to Pb. Section AX7.1.4 outlines 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 the more recent critical research conducted on 

effects of Pb on primary producers, consumers, and decomposers. 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 materials in controlled studies to permit evaluation 

of specific impacts on higher levels of organization (beyond the individual organism). Potential 

effects at the population level or higher are, of necessity, extrapolated from individual level 

studies. Available population, community, or ecosystem 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). Therefore, the best documented links 

between lead and effects on the environment are with effects on individual organisms. Impacts 

on terrestrial ecosystems are discussed 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 EPA to develop Ecological Soil Screening Levels 

(Eco-SSLs). Eco-SSLs describe the 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 by the U.S. EPA for use in screening-level assessments at 

Superfund sites to identify contaminants requiring 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 order to provide confidence that contaminants, 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 unlikely. Eco-SSLs were derived for terrestrial 

plants, soil invertebrates, birds, and mammals. Detailed procedures using an extensive list of 

acceptability and exclusion criteria (U.S. Environmental 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 the 
 AX7-59 

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 were incorporated into the development of the Eco-SSLs. 

For the plant and invertebrate Eco-SSLs, studies were scored to favor relatively high 

bioavailability. For wildlife Eco-SSLs, only species 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 varied diet, species were 

assumed to forage exclusively at the contaminated site, relative bioavailability or Pb in soil and 

diet was assumed to be 1, and the TRV was selected 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 studies, studies conducted 

on mine tailings, and studies conducted with hydroponic solutions); mixture toxicity (addressed 

in Section AX7.1.3); issues related to indirect effects (e.g., effects on predator/prey interactions, 

habitat alteration, etc.); and human health-related research (e.g., hypertension), which is 

addressed in other sections of this document. 

The toxicity data presented herein should be reviewed with a note of caution regarding 

their relevance to field conditions. Laboratory studies, particularly 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 environmental 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 vascular 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 Agency, 1986a) concluded that most plants 

experience reduced growth when Pb concentrations 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 experience 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 much 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 elongation (at <2.5 mg/kg), absence of root growth 

(at 5 mg/kg), or 55% inhibition of seed germination (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; effects were less severe when the Pb was added 

as Pb-oxide. 

Consumers 

The EPA (U.S. Environmental Protection Agency, 1986a) concluded that food is the 

largest contributor of Pb to animals, with inhalation rarely accounting for more than 10 to 15% 

of daily intake of Pb and drinking water exposures being quite low. It also was concluded that a 

regular dose of 2 to 8 mg/kg-day causes death in most animals. Grazing animals may consume 

more than 1 mg/kg-day in habitats near smelters 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. Environmental Protection Agency, 1986a). Lead may 

affect decomposition processes by direct toxicity 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 resistant 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 (depending 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 and presented an overall summary of 

conclusions and recommendations (U.S. Environmental Protection Agency, 1990). The major 

conclusion was that available laboratory and field data indicated that high concentrations of Pb 

can affect certain plants and alter the composition 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 Effects of Lead on Primary Producers 

Several studies published since 1986 have reported terrestrial plant exposure to Pb in soil, 

many of which were reviewed during the development of the Eco-SSLs (U.S. Environmental 

Protection Agency, 2005b). The relevant information from the Eco-SSL document (U.S. 

Environmental Protection Agency, 2005b) is summarized 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 photosynthetic and transpiration rates and decreased 

growth and yield (U.S. Environmental Protection Agency, 2005b). The phytotoxicity of Pb is 

considered relatively low, due to the limited availability and uptake of Pb from soil and soil 
 AX7-62 

solution and minimal translocation of Pb from roots to shoots (Påhlsson, 1989). Although many 

laboratory toxicity studies have reported effects on plants, there are few reports of phytotoxicity 

from Pb exposure under field conditions. For example, Leita et al. (1989) and Sieghardt (1990) 

reported high concentrations of Pb and other metals in soil and vegetation collected around 

mining areas in Europe, with no toxicity symptoms observed in plants or fruit. 

The literature search completed for the terrestrial plant Eco-SSL development identified 

439 papers for detailed review, of which 28 met the minimum criteria (U.S. Environmental 

Protection Agency, 2005a). Thirty ecotoxicological endpoints were gleaned from these 28 

papers and were further evaluated; most of those evaluated growth (biomass), which was 

considered the most sensitive and ecologically relevant endpoint (U.S. Environmental Protection 

Agency, 2005b). Five of the endpoints, representing 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 significant digits) (Table AX7-1.3.1). 

Table AX7-1.3.1. Plant Toxicity Data 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 Concentration, or the geometric mean of the NOAEC 

 (no-observed-adverse-effect concentration) and LOAEC (lowest-observed-adverse-effect concentration). 

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

The 25 ecotoxicological endpoints that were not used to develop the Eco-SSL for plants 

are presented in Table AX7-1.3.2. The first six endpoints 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 Concentration, or the geometric mean of the NOAEC 

 (no-observed-adverse-effect concentration) and LOAEC (lowest-observed-adverse-effect concentration). 

Source: U.S. Environmental Protection Agency (2005b). 
 AX7-64 

AX7.1.3.3 Recent Studies on the Effects of Lead on Consumers 

Since the 1986 Lead AQCD (U.S. Environmental 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 reviewed during development of the Eco-SSLs (U.S. 

Environmental Protection Agency, 2005b). The relevant information from the Eco-SSL 

document (U.S. Environmental Protection Agency, 2005b) is described below. A literature 

search and review was conducted to identify critical papers published since 2002. These recent 

critical papers are described briefly below. No studies were found that used inhalation exposures 

to evaluate endpoints such as survival, growth, and reproduction in birds or mammals. All 

studies described below exposed organisms via ingestion (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 simplified version of the equation used to 

calculate the Eco-SSL is: 

[( ) ( )]

TRV

C IR C IR BW

HQ soil× soil + food× food ÷ =

where: 

HQ = hazard quotient (1 mg Pb/kg bw/day) 

Csoil = concentration of Pb in soil (mg Pb/kg soil) 

IRsoil = incidental soil ingestion rate (kg soil/day) 

Cfood = concentration of Pb in food (mg Pb/kg food) 

IRfood = food ingestion rate (kg food/day) 

BW = body weight (kg) 

TRV = toxicity reference 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 influence on the resulting EcoSSL are the assumption of 100% bioavailability 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 different diets and calculating the Eco-SSL. The avian species 

selected were dove (herbivore), woodcock (insectivore), and hawk (carnivore). The mammalian 

species selected were vole (herbivore), 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 similar to those reported previously: mortality, 

changes in juvenile growth rate and weight gain, effects on various reproductive measures, and 

changes in behavior (U.S. Environmental Protection Agency, 2005b). Reproductive effects 

following Pb exposure included declines in clutch size, number of young hatched, and number of 

young fledged as well as decreased fertility or eggshell 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 species, of which 54 met the minimum criteria for 

further consideration for avian Eco-SSL development (U.S. Environmental Protection Agency, 

2005b). The 106 toxicological data points for birds 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 LOAEL (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. 

 
 AX7-66 Table AX7-1.3.3. Avian Toxicity Data Used to Develop the Eco-SSL Avian Species No. of Doses Route of Exposure Exposure Duration Duration Units Age Age Units Lifestage Sex Effect Type Effect Measure Response Site 

Reproduction Japanese quail 4 FD 5 w 6 w LB F REP PROG WO 

Chicken 3 FD 4 w NR NR LB F REP PROG WO Chicken 4 FD 30 d 22 w LB F EGG ALWT EG Mallard 2 FD 76 d NR NR SM F EGG ESTH EG American kestrel 3 FD 6 mo 1-6 yr AD F REP RSUC WO Japanese quail 5 FD 5 w 6 d JV M REP TEWT TE Japanese quail 5 FD 5 w 1 d JV M REP TEWT TE Japanese quail 3 FD 32 d NR NR AD F REP PROG WO Japanese quail 5 FD 12 w 0 d LB B REP EGPN EG Japanese quail 4 FD 12 w NR NR LB F REP PROG WO Chicken 5 FD 10 w NR NR LB F REP PROG WO Ringed turtle dove 2 DR 11 w NR NR AD M REP TEWT TE Japanese quail 2 FD 1 w 14 w JV F REP TPRD WO Japanese quail 

2 FD 27 d NR NR AD F REP PROG WO 
 AX7-67 Table AX7-1.3.3 (cont’d). Avian Toxicity Data Used to Develop the Eco-SSL Avian Species No. of Doses Route of Exposure Exposure Duration Duration Units Age Age Units Lifestage Sex Effect Type Effect Measure Response Site 

Growth Japanese quail 3 FD 5 w 1 d JV F GRO BDWT WO Japanese quail 3 FD 2 w 1 d JV B GRO BDWT WO Japanese quail 2 FD 2 w 1 d JV NR GRO BDWT WO Japanese quail 3 FD 4 w 0 d JV F GRO BDWT WO Chicken 4 FD 4 w 4 w JV NR GRO BDWT WO Chicken 4 FD 4 w 4 w JV NR GRO BDWT WO Japanese quail 5 FD 12 w 0 d JV F GRO BDWT WO Japanese quail 5 FD 12 w 1 w JV F GRO BDWT WO Japanese quail 5 FD 2 w 6 d JV NR GRO BDWT WO Japanese quail 5 FD 1 w 1 d JV NR GRO BDWT WO Chicken 2 FD 21 d 1 d JV B GRO BDWT WO Duck 3 GV 3 mo 24 w MA F GRO BDWT WO American kestrel 4 GV 10 d 1 d JV NR GRO BDWT WO Chicken 2 FD 20 d 1 d JV B GRO BDWT WO Japanese quail 

5 FD 14 d 1 d JV B GRO BDWT WO 
 AX7-68 Table AX7-1.3.3 (cont’d). Avian Toxicity Data Used to Develop the Eco-SSL Avian Species No. of Doses Route of Exposure Exposure Duration Duration Units Age Age Units Lifestage Sex Effect Type Effect Measure Response Site 

American kestrel 4 FD 60 d 1-2 yr AD B GRO BDWT WO Chicken 5 FD 2 w 1 d JV M GRO BDWT WO Mallard 4 FD 8 d 9 d JV NR GRO BDWT WO Chicken 5 FD 20 d 1 d JV M GRO BDWT WO Chicken 2 FD 3 w 1 d JV M GRO BDWT WO Japanese quail 3 FD 32 d NR NR AD F GRO BDWT WO Chicken 2 FD 19 d 1 d JV M GRO BDWT WO Chicken 3 FD 2 w 1 d JV M GRO BDWT WO Chicken 4 FD 14 d 8 d JV M GRO BDWT WO Chicken 2 FD 20 d 1 d JV M GRO BDWT WO Chicken 2 OR 4 w NR NR JV B GRO BDWT WO Chicken 2 FD 7 d 1 d JV M GRO BDWT WO Chicken 2 FD 2 w 1 d JV M GRO BDWT WO Chicken 2 FD 7 d 1 d IM NR GRO BDWT WO Chicken 2 FD 14 d 8 d JV M GRO BDWT WO AD = adult; ALWT = albumin weight; B = both; BDWT = body weight changes; d = days; DR = drinking water; EG = egg; EGG = effects on eggs; EGPN = egg produ

thinning; F = female; FD = food; GRO = growth; GV = gavage; JV = juvenile; LB = laying bird; MA = mature; M = male; mo = months; NR = not reported; OR = otheor numbers; REP = reproduction; RSUC = reproductive success; SM = sexually mature; TE = testes; TEWT = testes weight; TPRD = total production; w = weeks; WO 

Source: U.S. Environmental Protection Agency (2005b) 
 AX7-69 

Figure AX7-1.3.1. Avian reproduction and growth 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 Protection Agency, 2005b) recognizes that toxicity is observed over a 

wide range of doses (<1 to >100 mg Pb/kg bw/day), even when considering only reproductive 

effects in the same species. In addition, the TRV 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, behavioral, physiological, and 

pathological effects, which generally are observed at lower 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, several studies have been conducted on waterfowl. 

Toxicity data for waterfowl (in particular, mallards) were included in the soil Eco-SSL 

development process (Table AX7-1.3.3), although 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 strategy of the species (which may vary during 

reproduction), and the nutritional status of the animal. Sediment is recognized as an important 

route of exposure for waterfowl, particularly those species that dabble (i.e., forage on 

invertebrates in the sediment) (Beyer et al., 2000; Douglas-Stroebel et al., 2005). DouglasStroebel et al. (2005) found that mallard ducklings exposed to Pb-contaminated sediment and a 

low nutrition diet exhibited more changes in behavior (as measured by time bathing, resting, and 

feeding) than Pb exposure or low nutrition exposure alone. These effects may be due to the low 

nutrition diet being deficient in levels of protein, amino acids, calcium, zinc, and other nutrients. 

Beyer et al. (2000) related blood Pb to sublethal effects in waterfowl along the Coeur 

d=Alene River near a mining site in Idaho. The authors suggested that 0.20 mg/kg ww blood Pb 

represents the no-effect level. This no-effect blood concentration corresponds to a sediment Pb 

concentration of 24 mg/kg. A sediment concentration 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 Scheuhammer (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 concluded that some swan mortality 

may occur at blood Pb levels of 1.9 mg/kg ww, corresponding 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 consuming sediment at the 90th percentile rate 

would die with chronic exposure to sediment 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 similar to those reported previously: mortality, 

effects on reproduction, developmental effects, and changes in growth (U.S. Environmental 
 AX7-71 

Protection Agency, 2005b). Very little research has been done to expand the knowledge of the 

toxicity of Pb to mammalian wildlife, since 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 for Eco-SSL development, 219 met 

the minimum criteria for further consideration for mammalian Eco-SSL development (U.S. 

Environmental Protection Agency, 2005b). The 343 ecotoxicological endpoints for mammals 

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 AX7-1.3.4, Figure AX7-1.3.2). The geometric mean of the NOAELs was 

calculated as 40.7 mg/kg-day, which was higher than the lowest bounded LOAEL for survival, 

growth, or reproduction. Therefore, the highest bounded NOAEL that was lower than the lowest 

bounded LOAEL for survival, growth, or reproduction (4.7 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 56 mg/kg soil (U.S. Environmental 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 Eco-SSL document (U.S. Environmental Protection 

Agency, 2005b) reveals that effects on survival generally are observed at Pb doses much greater 

than those reported in the 1986 Lead AQCD, where 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 other studies reported no mortality or a LOAEL for mortality at 

significantly higher doses (U.S. Environmental Protection Agency, 2005b). The five studies that 

supported this low toxic level were conducted on three species (mouse, rat, and cow) and used 

either gavage or drinking water as the exposure method. The 34 other studies included data on 

these three species as well as five other species (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-day (U.S. Environmental Protection Agency, 

2005b). Therefore, the review of data in the Eco-SSL document suggests effects on survival of 
 AX7-72 Table AX7-1.3.4. Mammalian Toxicity Data Used to Develop the Eco-SSL Mammalian Species No. of Doses Route of Exposure Exposure Duration Duration Units Age Age Units Lifestage Sex Effect Type Effect Measure Respon

Site 

Reproduction Rat 5 DR 62 d 21 d GE F REP PRWT WO Rat 6 DR 21 d NR NR GE F REP PRWT WO Rat 3 DR 35 d NR NR AD M REP RSUC WO Rat 4 DR 62 d 21 d GE B REP PRWT WO Sheep 3 FD 27 w NR NR GE F REP RSUC WO Rat 6 DR 21 d NR NR GE F REP PRWT WO Guinea pig 3 DR 40 d NR NR GE F REP PRWT WO Rat 5 FD 92 w 21 d JV M REP TEWT TE Rat 4 DR 23.8 d 21 d LC F REP Other WO Rat 5 DR 23.8 d 21 d GE F REP Other WO Cotton rat 3 DR 7 w NR NR AD M REP RHIS RT Rat 4 GV 9 w 10 w JV M REP SPCV TE Rat 3 DR 100 d 21 d GE F REP PRWT WO Rat 2 FD 35 d 70 d LC F REP PRWT WO Rat 4 DR 60 d NR NR SM M REP TEWT TE Rat 4 DR 56 d 70 d LC F REP PROG WO Rat 3 DR 31 d NR NR LC F REP PRWT WO Rat 4 GV 41 d NR NR GE F REP PRWT WO Rat 5 DR 1 w 94 d JV M REP SPCL SM Rat 4 DR 30 d NR NR SM M REP Other SV 
 AX7-73 Table AX7-1.3.4 (cont’d). Mammalian Toxicity Data Used to Develop the Eco-SSL Mammalian Species No. of Doses Route of Exposure Exposure Duration Duration Units Age Age Units Lifestage Sex Effect Type Effect Measure Respon

Site 

Hamster 2 DR 51 d 15 w GE F REP PROG WO Hamster 2 DR 14 d 11 w GE F REP PROG WO Rat 4 DR 37 d NR NR GE F REP PRWT WO Rat 5 GV 12 d NR NR GE F REP RSEM EM Rat 3 DR 68 d 25 d GE F REP PRWT WO Rat 4 DR 77 d 25 d GE F REP PRWT WO Rat 2 DR 21 d NR NR LC F REP PRWT WO Mouse 3 FD 8 w 2 mo GE M REP SPCV TE Mouse 7 FD 30 d NR NR LC F REP PRWT WO Mouse 7 FD 30 d NR NR LC F REP PRWT WO Rat 4 DR 21 d NR NR GE F REP DEYO WO Rat 5 DR 10 w NR NR AD M REP TEWT MT Rat 2 GV 102 d 30 d GE F REP PRWT WO Rat 2 DR 9 mo NR NR SM M REP RHIS TE Rat 4 FD 4 d NR NR LC F REP PRWT WO Rat 4 DR 13 w NR NR JV M REP FERT WO Mouse 4 GV 60 d NR NR AD F REP RPRD OV Rat 3 FD 339 d 26-27 d JV B REP PRWT WO Rat 2 DR 9 mo 21 d JV F REP DEYO WO Mouse 2 DR 6 mo 21 d JV F REP DEYO WO Mouse 4 GV 52 d 2 mo GE F REP PROG EM Rat 2 DR 120 d 1 d GE M REP SPCL TE 
 AX7-74 Table AX7-1.3.4 (cont’d). Mammalian Toxicity Data Used to Develop the Eco-SSL Mammalian Species No. of Doses Route of Exposure Exposure Duration Duration Units Age Age Units Lifestage Sex Effect Type Effect Measure Respon

Site 

Mouse 2 DR 5 w NR NR AD M REP TEDG TE Mouse 2 GV 2 w NR NR JV M REP SPCL SM Rat 2 FD 102 d NR NR GE F REP PROG WO Rat 2 GV 3 mo 8 w SM M REP TEDG TE Rat 2 DR 18 d NR NR GE F REP PRWT WO Rat 3 DR 90 d NR NR AD M REP SPCL SM Mouse 2 DR 23 d NR NR GE F REP PRWT WO Mouse 2 DR 62 d NR NR GE F REP PRWT WO Mouse 2 DR 18 w 6-8 w LC F REP PRWT WO Mouse 2 DR 12 w 9 w SM M REP PRFM WO Mouse 4 FD 18 d NR NR GE F REP PRWT WO Rat 2 DR 4 w 99 d JV M REP SPCL SM Rat 5 DR 6 w 4 mo GE F REP RHIS WO Rat 2 DR 22 d NR NR GE F REP PRWT WO Rat 3 DR 30 d 52 d JV M REP GREP PG Rat 2 DR 13 w NR NR GE F REP PRWT WO Rat 2 DR 21 d 80 d JV F REP PRWT WO Rat 4 FD 3 w NR NR LC F REP PRWT WO Rat 2 FD 1 w 19 w LC F REP PRWT WO Rat 4 FD 3 w NR NR LC F REP PRWT WO Rat 3 FD 25 d NR NR LC F REP PRWT WO Rat 2 DR 21 d NR NR LC F REP PRWT WO 
 AX7-75 Table AX7-1.3.4 (cont’d). Mammalian Toxicity Data Used to Develop the Eco-SSL Mammalian Species No. of Doses Route of Exposure Exposure Duration Duration Units Age Age Units Lifestage Sex Effect Type Effect Measure Respon

Site 

Rat 2 DR 30 d 52 d JV M REP SPCL SM Rat 2 DR 17 d NR NR GE F REP PRWT WO Rat 2 DR 24 d NR NR LC F REP PRWT WO Rat 2 DR 12 d NR NR GE F REP PRWT WO Rat 2 DR 30 d 27 d JV M REP SPCL SM Mouse 2 DR 44 d NR NR GE F REP PRWT WO Mouse 2 DR 14 d NR NR LC F REP PRWT WO Rat 2 DR 50 d 24 d JV F REP RBEH WO Mouse 2 DR 45 d 50-100 d GE F REP ODVP WO Rat 2 DR 22 d NR NR GE F REP PRWT WO Mouse 2 DR 48 d NR NR GE F REP PRWT WO Rat 2 DR 9 mo 3 mo SM M REP SPCL TE Rat 2 DR 9 mo NR NR SM M REP TEDG TE Rat 2 DR 3 w 14 w LC F REP PRWT WO Mouse 2 FD 7 d NR NR GE F REP RSUC EM Rat 2 DR 126 d 1 d GE F REP PROG WO Rat 2 DR 20 w 10 w GE F REP PRWT WO Mouse 2 DR 4 d NR NR LC F REP PRWT WO Rat 2 FD 2 w NR NR LC F REP PRWT WO Rat 2 FD 7 d NR NR LC F REP PRWT WO Rat 2 FD 21 d NR NR LC F REP PRWT WO Mouse 4 DR 18 w 11 w JV F REP TEWT WO 
 AX7-76 Table AX7-1.3.4 (cont’d). Mammalian Toxicity Data Used to Develop the Eco-SSL Mammalian Species No. of Doses Route of Exposure Exposure Duration Duration Units Age Age Units Lifestage Sex Effect Type Effect Measure Respon

Site 

Rat 2 FD 30 d NR NR LC F REP PRWT WO Mouse 2 DR 14 w NR NR GE B REP PROG WO Rat 2 FD 16 d NR NR LC F REP PROG WO Rat 2 FD 7 d NR NR LC F REP PRWT WO Rat 2 FD 25 d NR NR LC F REP PRWT WO Rat M FD 27 d NR NR LC C REP PROG WO Mouse 2 DR 14 w 21 d JV B REP PROG WO Rat 2 FD 17 d NR NR LC F REP PRWT WO Growth Horse 2 FD 15 w 20-21 w JV M GRO BDWT WO Rat 2 FD 21 d 0 d JV F GRO BDWT WO Rat 6 DR 21 d NR NR GE F GRO BDWT WO Rat 5 DR 7 d 50 d AD F GRO BDWT WO Cattle 4 OR 7 w 1 w JV M GRO BDWT WO Rat 3 DR 14 d 21 d JV F GRO BDWT WO Rat 2 DR 332 d 28 d JV B GRO BDWT WO Rat 4 DR 7 w 21 d GE F GRO BDWT WO Dog 3 FD 7 mo NR NR JV NR GRO BDWT WO Rat 3 DR 30 d 22-24 d JV M GRO BDWT WO Rat 4 DR 23 d 22 d JV F GRO BDWT WO Cattle 3 OR 84 d NR NR JV M GRO BDWT WO Rat 2 OR 6 w NR NR AD M GRO BDWT WO 
 AX7-77 Table AX7-1.3.4 (cont’d). Mammalian Toxicity Data Used to Develop the Eco-SSL Mammalian Species No. of Doses Route of Exposure Exposure Duration Duration Units Age Age Units Lifestage Sex Effect Type Effect Measure Respon

Site 

Rat 2 GV 8 w NR NR JV F GRO BDWT WO Rat 3 DR 6 mo NR NR AD M GRO BDWT WO Rabbit 3 GV 10 d 1 d JV F GRO BDWT WO Rat 2 DR 140 d 21 d JV M GRO BDWT WO Rat 2 DR 6 w NR NR JV M GRO BDWT WO Rat 2 FD 10 w NR NR JV M GRO BDWT WO Rat 2 OR 6 w NR NR AD M GRO BDWT WO Rat 2 DR 7 w NR NR JV M GRO BDWT WO Mouse 3 DR 14 d 0 d JV NR GRO BDWT WO Rat 4 GV 9 w 10 w JV M GRO BDWT WO Rat 3 FD 339 d 26-27 d JV B GRO BDWT WO Rat 4 GV 29 d NR NR SM F GRO BDWT WO Rat 7 DR 10 w NR NR JV M GRO BDWT WO Rat 4 DR 56 d 70 d LC F GRO BDWT WO Sheep 5 FD 84 d NR NR JV M GRO BDWT WO Rat 2 DR 10 w NR NR JV M GRO BDWT WO Cattle 4 FD 7 w 16 w JV M GRO BDWT WO Rat 2 GV 28 d 2 d JV B GRO BDWT WO Rat 5 DR 4 w 94 d JV M GRO BDWT WO Rat 4 GV 12 d 2 d JV B GRO BDWT WO Rat 2 FD 4 w NR NR JV M GRO BDWT WO Rat 6 DR 10 w NR NR JV M GRO BDWT WO 
 AX7-78 Table AX7-1.3.4 (cont’d). Mammalian Toxicity Data Used to Develop the Eco-SSL Mammalian Species No. of Doses Route of Exposure Exposure Duration Duration Units Age Age Units Lifestage Sex Effect Type Effect Measure Respon

Site 

Mouse 3 FD 4 w 3 mo JV B GRO BDWT WO Mouse 2 DR 18 w 6-8 w LC F GRO BDWT WO Mouse 2 DR 12 w NR NR GE M GRO BDWT WO Rat 3 DR 30 d 52 d JV M GRO BDWT WO Rat 2 DR 4 w 99 d JV B GRO BDWT WO Rat 4 GV 18 d 3 d JV M GRO BDWT WO Mouse 3 DR 6 w 7 w SM M GRO BDWT WO Rat 4 GV 18 d 2 d JV B GRO BDWT WO Rat 2 GV 91 d NR NR JV M GRO BDWT WO Rat 2 DR 21 d 80 d JV F GRO BDWT WO Rat 4 FD 1 w NR NR LC F GRO BDWT WO Rat 4 DR 30 d NR NR JV M GRO BDWT WO Mouse 5 DR 10 w NR NR JV M GRO BDWT WO Rat 2 DR 30 d 52 d JV M GRO BDWT WO Rat 4 GV 14 d 14 d JV NR GRO BDWT WO Rat 5 GV 14 d 20 d JV NR GRO BDWT WO Rat 2 FD 14 mo 0 d JV NR GRO BDWT WO Rat 2 GV 102 d 30 d LC F GRO BDWT WO Mouse 4 GV 12 d 6 d JV M GRO BDWT WO Mouse 7 FD 30 d NR NR LC F GRO BDWT WO Rat 2 DR 126 d 1 d GE F GRO BDWT WO Rat 2 DR 20 w 10 w GE F GRO BDWT WO 
 AX7-79 Table AX7-1.3.4 (cont’d). Mammalian Toxicity Data Used to Develop the Eco-SSL Mammalian Species No. of Doses Route of Exposure Exposure Duration Duration Units Age Age Units Lifestage Sex Effect Type Effect Measure Respon

Site 

Mouse 7 FD 28 d NR NR LC F GRO BDWT WO Rat 4 FD 18 d NR NR LC F GRO BDWT WO Rat 2 DR 9 d 21 d JV M GRO BDWT WO Cattle 2 FD 283 d 7 mo JV M GRO BDWT WO Rat 3 DR 92 d 25 d GE F MPH GMPH TB Rat 4 DR 7 d 25 d GE F GRO BDWT WO Rat 2 DR 5 d 26 d JV F GRO BDWT WO Rat 2 DR 26 d 22 d JV F GRO BDWT WO Rat 2 DR 14 d 26 d JV F MPH Other TA Rat 2 DR 10 d 26 d JV F GRO BDWT WO Mouse 2 GV 3 w NR NR JV M GRO BDWT WO Dog 2 OR 5 w <1 yr JV NR GRO BDWT WO Shrew 4 FD 31 d NR NR JV B GRO BDWT WO Rat 3 GV 58 d 2 d JV B GRO BDWT WO Pig 2 FD 13 w 4 w JV NR GRO BDWT WO Rat 2 GV 29 d 2 d JV F GRO BDWT WO Rat 2 FD 5 w NR NR MA NR GRO BDWT WO Rat 2 GV 6 d 1 d JV B GRO BDWT WO Rat 2 DR 30 d 27 d JV M GRO BDWT WO Rat 2 DR 50 d 24 d JV M GRO BDWT WO Rat 2 GV 28 d 2 d JV M GRO BDWT WO Rat 4 GV 14 d 18 d JV NR GRO BDWT WO 
 AX7-80 Table AX7-1.3.4 (cont’d). Mammalian Toxicity Data Used to Develop the Eco-SSL Mammalia n Species No. of Doses Route of Exposure Exposure Duration Duration Units Age Age Units Lifestage Sex Effec t Type Effect Measur e Respon

Site 

Mouse 2 DR 45 d 50-100 d GE F GRO BDWT WO Rat 4 FD 1 w NR NR LC F GRO BDWT WO Rat 2 DR 6 w 14 w LC F GRO BDWT WO Mouse 4 DR 10 w 11 w JV F GRO BDWT WO Rat 2 FD 21 d NR NR LC F GRO BDWT WO Rat 2 GV 18 d 2 d JV B GRO BDWT WO Rat 2 FD 2 w 0 d JV NR GRO BDWT WO Rat 4 GV 14 d 24 d JV NR GRO BDWT WO Rat 2 FD 2 w 60-80 d JV M GRO BDWT WO Rat 3 GV 14 d 16 d JV NR GRO BDWT WO Rat 2 FD 14 d 60 d JV M GRO BDWT WO AD = adult; B = both; BDWT = body weight changes; d = days; DEYO = death of young; DR = drinking water; F = female; FD = food; FERT = fertility; GMPH = gen

GRO = growth; GV = gavage; JV = juvenile; LC = lactation; M = male; MA = mature; mo = months; MPH = morphology; NR = not reported; ODVP = offspring deve

PG = prostate gland; PROG = progeny counts or numbers; PRWT = progeny weight; RBPH = reproductive behavior; REP = reproduction; RHIS = reproductive organ 

embryos; RSUC = reproductive success (general); RT = reproductive tissue; SM = sexually mature; SPCL = sperm cell counts; SPCV = sperm cell viability; TA = tail; 

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 Effects 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 earthworms 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. Many of these were reviewed during the development 

of the Eco-SSLs (U.S. Environmental Protection 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 invertebrates observed in studies conducted since the 

1986 Lead AQCD are similar to those reported previously: 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 terrestrial invertebrate Eco-SSL development 

identified 179 papers for detailed review, of which 13 met the minimum criteria for further 

consideration (U.S. Environmental Protection Agency, 2005b). Most of the 18 ecotoxicological 

endpoints that were further evaluated measured reproduction or survival as the ecologically 

relevant endpoint. Four of these, representing one species under three different pH test 

conditions were used to develop 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 MATC1 (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) and LOEC (lowest-observed-effect concentration). 

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

In a study designed to test the toxicity of Pb to the earthworm Eisenia fetida, Davies 

et al. (2002) found that the 28-day LC50 (± 95% confidence intervals) for Pb in soils 

contaminated with Pb(NO3)2 was 4379 ∀ 356 mg/kg. Twenty-eight day EC50 values 

(±95% confidence intervals) for weight change and cocoon production were 1408 ∀ 198 and 
 AX7-83 

971 ∀ 633 mg/kg, respectively. Significant mortalities were noted at concentrations of 2000 

mg/kg. These data are consistent with those reported in the Eco-SSL document (U.S. 

Environmental Protection Agency, 2005b) for the same species of earthworm. 

Nursita et al. (2005) found no mortality and 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) determined a NOEC (no-observed-effect 

concentration) and LOEC (lowest-observed-effect concentration) for collembolan reproduction 

of 2000 and 5000 mg/kg, respectively. (A MATC [maximum-acceptable-threshold 

concentration] of 3162 mg/kg was used to develop the Eco-SSL). 

The remaining 14 toxicity endpoints that were not used to develop the Eco-SSL for 

invertebrates are presented in Table AX7-1.3.6. None of these endpoints was considered eligible 

for Eco-SSL derivation. 

Lock and Janssen (2002) exposed the potworm Enchytraeus albidus to Pb, as Pb-nitrate. 

The 21-day LC50 was 4530 mg/kg, and the 42-day EC50 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 EC50 for the F1 generation (394 mg/kg) was 

similar to that of the P generation. The authors 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 LC50 reported for 

the potworm was higher than reported for nematodes and similar to that reported for the 

earthworm. The EC50 for reproduction was lower than reported for the earthworm or collembola. 

Recent work by Bongers et al. (2004) cautioned against 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 toxicity of the counterion (nitrate) was not taken 

into account during Eco-SSL development. Percolation (removal of the counterion) had no 

statistically significant effect on Pb-chloride toxicity (LC50 = 2900 mg/kg for both nonpercolated and percolated soil; EC50 for reproduction = 1900 mg/kg or 2400 mg/kg for 
 AX7-84 

Table AX7-1.3.6. Invertebrate Toxicity Data 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 LC50 285 

Nematode 4 1.14 LC50 297 

Nematode 4 4.2 LC50 847 

Nematode 4 4.2 LC50 1341 

Nematode 6.2 1.7 LC50 1554 

Nematode 5.1 3.0 LC50 891 

Earthworm 6.3 10.0 EC50 1940 

Earthworm 6.1 10.0 EC50 1629 

Earthworm 6.0 10.0 LC50 3716 

Earthworm 6.5 10.0 ILL 1.16 

Nematode 4 10 LC50 1434 

Nematode 4 10 NOAEC 2235 

Nematode 6.1 3.4 LC50 13.9 

Nematode 6.2 2.2 LC50 11.6 

*NOAEC (no-observed-adverse-effect concentration); LC50 (concentration lethal to 50% of test population); 

 EC50 (effect concentration for 50% of test population); 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 (LC50 = 980 mg/kg or 2200 mg/kg for non-percolated or percolated 

soil, respectively; EC50 for reproduction = 580 mg/kg or 1700 mg/kg for non-percolated or 

percolated soil, respectively). Lead nitrate was more toxic than Pb-chloride for survival and 

reproduction. However, the toxicity of Pb, from chloride or nitrate, was not significantly 

different after the counterion was percolated 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 in 
 AX7-85 

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 LC50 concentration of 11 mg/kg. Exposure for 14 days resulted in a number 

of effects including body fragmentation, protrusions, rupture of the cuticle, etc. Many of these 

effects may trigger defensive mechanisms. For example, fragmentation of the affected posterior 

region was followed by regeneration and a new ectoderm layer was formed to cover affected 

areas, both of which processes may serve to prevent 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 Directive 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 unacceptable risk to microbes and not be reflected 

as unacceptable risks to higher organisms; that the significance of laboratory-derived effects data 

to the ecosystem is uncertain; and that the spatial (across millimeter distances) and temporal 

(within minutes to hours) variation makes understanding ecological consequences challenging. 

Microbial endpoints often vary dramatically based on moisture, temperature, oxygen, and many 

non-contaminant factors. Therefore, 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 but that they should be addressed 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 populations are difficult to determine and interpret, 

indicators for soil communities are often measured as proxies for toxicity (e.g., urease activity in 

soil). Recent studies of this nature (Doelman and Haanstra, 1986; Wilke, 1989; Haanstra and 

Doelman, 1991) are summarized in this subsection. The Pb concentrations in these recent 

studies (1000 to 5000 mg/kg) are consistent with those reported in the 1986 Lead AQCD (U.S. 

Environmental Protection Agency, 1986a) as associated with effects on microbial processes 

(750 to 7500 mg/kg). 
 AX7-86 

The effects of Pb-chloride on the processes of nitrification and nitrogen mineralization 

were studied in a 28-day experiment by Wilke (1989). 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 (Haanstra and Doelman, 1991) and urease activity 

(Doelman and Haanstra, 1986) in soil were investigated. LC50s 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 LC50 for a decrease in urease activity 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 include decreased photosynthetic and transpiration 

rates in addition to decreased growth 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 Agency, 2005b) are lower than those discussed in 

the 1986 Lead AQCD, although both documents acknowledged 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, organic 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 agricultural plant species (i.e., 
 AX7-87 

vegetable and grain crops). Few data are available for trees or native 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 Protection 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 derived for other species would contribute to the 

understanding of Pb toxicity, particularly for wildlife species with different gut physiologies. 

In addition, data derived using environmentally realistic exposures, such as from Pbcontaminated soil and food, may be recommended. Finally, data derived from inhalation 

exposures, which evaluate endpoints such as survival, growth, and reproduction, would 

contribute to understanding the implications 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 mineralization and enzyme activities. 

Recent data on 1986 Lead toxicity to soil invertebrates and microorganisms are consistent with 

those reported in the 1986 Lead 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 significantly 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 subjected 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 ingestion 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 Natural Terrestrial Ecosystems 

The concept that organisms are part of larger systems that include both biotic and abiotic 

components of the environment dates back to the naturalists of the Victorian era. However, the 

breakthrough in what we now consider the ecosystem approach to ecology occurred in the 1950s 

and 1960s when E.P. and H.T. Odum pioneered the quantitative analysis of ecosystems 

(Odum, 1971). This approach encouraged the calculation of energy flows into, out of, and 

within explicitly defined ecosystems. The rapid development of computer 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 quantitative analysis of ecosystems was extended to examine 

the flows of nutrients and other chemical compounds. 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 ecosystem. In a system 

limited by nitrogen, for example, elevated concentrations of NO3

!

 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 study of trace element 

biogeochemistry. Input-output budgets can be used to determine whether an ecosystem is a net 

source or sink of a trace element. Changes to the input-output balance over time can be used to 

assess the effects of natural or experimental changes in deposition, land use, climate, or other 
 AX7-89 

factors. In addition, examination of fluxes within the ecosystem (in plant uptake, soil solutions, 

etc.) can be used to understand the processes that are most influential 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 misleading (Likens, 1989). To nurture long-term 

research, the National Science Foundation supports a network of Long-Term Ecological 

Research (LTER) sites that represent various biomes. 

This section describes terrestrial ecosystem research on Pb, focusing on work done since 

the 1986 Lead AQCD (U.S. Environmental Protection Agency, 1986a) and highlighting key 

long-term studies. Unfortunately, there are few studies that feature long-term data on trace metal 

behavior at multiple levels of organization. Therefore, this examination 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 Ecosystem 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, reductions 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 vigor or growth of vegetation results in a lower uptake of nitrogen 

and other nutrients from soils. Reduced biodiversity 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 obvious. If the stress reduces microbial activity rates, then 

nutrients bound in soil organic matter (e.g., organic nitrogen compounds) will likely be 

mineralized at a lower rate and retained 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; Likens 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 terrestrial ecosystems is strongly related to the 

organic matter cycle (Section AX7.1.2), stressors that could lead to disruption or alteration of the 

soil organic matter pool are of particular concern in assessing effects of ecosystem stress on Pb 

cycling. By binding soluble Pb, soil organic matter acts as a barrier 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 dissolve organic matter (DOM). This release 

of soluble Pb does not typically result in elevated surface water 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 the organic matter pool is dissolved at any time; 

(2) DOM-Pb complexes solubilized in upper soil horizons 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 ecosystems. 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 soils 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 resulting 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 the Rothamsted Experiment Station in England, 

long-term (i.e., >100 years) soil acidification significantly increased the mobility of Pb in the soil 
 AX7-91 

(Blake and Goulding, 2002). However, the increased mobility was only observed in very acid 

soils, those with pH of <4.5. The fraction of exchangeable Pb (extracted with 0.1 M CaCl2) 

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. Similarly, Nouri and Reddy (1995) observed higher levels of 

diethylenetriaminepentaacetic-acid-[DTPA]-extractable Pb in soils in a loblolly pine forest 

treated with simulated acid rain, but only in the most acidic treatment, with simulated rain with a 

pH of 3.5. 

Although acidification may increase 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 waters and concluded that Pb concentrations were governed by the 

DOM concentration, which masked any association with acidification. In an in situ lysimeter 

study, Bergkvist (1986) measured lower concentrations of Pb in soil solutions draining 

experimentally acidified plots than in unacidified plots. In a laboratory study using large soil 

columns, Merino and García-Rodeja (1997) observed no effect of experimental acidification on 

the release of Pb to soil solution. Thus, while acidification may increase the 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 drainage 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 Johnson, 1980; Turner et al., 1985). If acidification 

results in the mobilization of Pb and organic matter into these mineral soils, then increased 

streamwater Pb concentrations would likely follow. 
 AX7-92 

Land Use and Industry 

Changes in land use also represent potentially significant changes in the cycling of 

organic matter in terrestrial ecosystems. Conversion of pasture and croplands to woodlands 

changes the nature and quantity of organic matter inputs to the soil. In temperate 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 trace 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 fractionation, despite significant 

acidification of the soils. Afforestation had 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 activity 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 declines in organic carbon in surface (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 acetate 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 organic carbon. This suggests that Pb mobilized during 

organic matter decomposition can resorb to remaining 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 disruption 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 results 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 temperatures. Together, the higher 

moisture and temperature in surface soils tend 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 after clear-cutting (Covington, 1981; Federer, 1984; Johnson et al., 

1995b). This loss of organic matter from the forest 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 very small increases in Pb concentrations in soil 

solutions draining the Oa soil horizon despite substantial reductions in the organic matter mass of 

that horizon (Fuller et al., 1988; Johnson et al., 1995b). These increases were associated with 

similarly small increases in dissolved organic carbon (DOC) concentrations 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 decreases in the Pb content of the Oa horizon 

(“H” in the European system of soil classification) 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 mineral soil horizons. The Pb decline in the Oa horizon 

was accompanied by a decrease in the organic matter content, leading the authors to attribute the 

Pb dynamics to leaching with DOM. In a study 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 watershed, which were uncut. The DOC and H+

 outputs 

were also lower in the clear-cut area. These patterns persisted in all 5 years of the study. 

Forest harvesting is a severe form of ecosystem disturbance, and, thus, it is somewhat 

surprising that studies of clear-cutting have shown little or no effect on Pb mobility or loss from 

forest ecosystems. Perhaps the strong complexation behavior of Pb with natural organic matter 

results in the retention of Pb in forest soils. Even in cases where Pb is mobilized in forest floor 

soils (Fuller et al., 1988; Berthelsen and Steinnes, 1995), there is no evidence of loss of Pb from 

the ecosystem, indicating that mineral soils are efficient in capturing and retaining any Pb that is 

mobilized in the forest floor. Therefore, the principal risk associated with forest harvesting is the 

loss of Pb in particulate form to drainage waters through erosion. In some relatively remote 
 AX7-94 

lakes in the United Kingdom the distribution of sediment-bound trace elements (including Pb) 

have been affected by forestry activities and 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 soil in-wash could dilute sediment heavy metal 

levels. 

Climate Change 

Atmospheric Pb is not likely to contribute significantly to global climate change. Lead 

compounds have relatively short residence times in the atmosphere, making it unlikely that they 

will reach the stratosphere. Also, Pb compounds are not known to absorb infrared radiation and, 

therefore, are unlikely to contribute to stratospheric ozone depletion or global warming. 

Climate change does, however, represent a disturbance to terrestrial ecosystems. 

Unfortunately, the potential linkages between climate-related stress 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 increase in temperature leading to increased rates 

of organic matter decomposition could lead to temporary 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 fluctuations of temperature and/or precipitation in 

terrestrial ecosystems. For example, there is some evidence for recent increases in the frequency 

of soil freezing events in the northeastern United States (Mitchell 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 horizons 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 contamination of floodplain ecosystems. Soils 

collected from the floodplain of the Elbe River, in Germany, contained elevated concentrations 

of Pb and other trace metals (Krηger and Grongroft, 2003). Tissues of plants from floodplain 

sites did not, however, contain higher Pb concentrations than control sites. More frequent 
 AX7-95 

or more severe flooding would likely result in increased inputs of Pb and other metals to 

floodplain soils. 

AX7.1.4.2 Effects of Lead Exposure on Natural Ecosystem Structure and Function 

The effects of Pb exposure on natural ecosystems are confounded by the fact that Pb 

exposure cannot be decoupled from 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 industrial activities are accompanied by other trace 

metals (e.g., Zn, Cu, and Cd) and sulfur dioxide (SO2) that may cause toxic effects independently 

or in concert with Pb. Reductions in the use of alkyl-Pb additives in gasoline have resulted in 

significant decreases in Pb deposition 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 significant reductions in the acidity (i.e., increased pH) of precipitation 

in the region (Likens et al., 1996; Driscoll et al., 1998). Therefore, changes in ecosystem Pb 

fluxes may be the result of reduced Pb inputs and/or reduced acidity. 

Experimental manipulation studies do not suffer from these confounding effects, because 

Pb can be added in specific amounts, with or without other compounds. Unfortunately, 

ecosystem-level manipulations involving Pb additions have not been undertaken. Therefore, we 

must use observations from field studies of Pb behavior in sites exposed to various forms of Pb 

pollution to assess the effects of Pb on terrestrial 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 cycling 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 related 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 ecosystems 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 enriched in Pb, arsenic, copper, cadmium, and 
 AX7-96 

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 soils 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 sites retained functioning forest stands. 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 affected sites were classified as bare ground. The 

occurrence of bare ground was significantly correlated with the phytotoxicity scores derived by 

Kapustka et al. (1995), indicating a link between phytotoxicity and the 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 decreased, 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 Anaconda 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 area 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 Palmerton (Ketterer et al., 2001), which lies on the 

north side of a gap in Blue Mountain, a ridge running roughly east-west 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 surrounding natural landscape is predominantly oak 

forest (Sopper, 1989; Storm et al., 1994). Biodiversity in affected areas is considerably lower 

than at reference sites, a pattern attributed to emissions from the smelters (Beyer et al., 1985; 
 AX7-97 

Sopper, 1989). The history is complicated, however, by the land use history of the area. 

Logging and fire in the early 20th century may also have played a role in the changes in the 

terrestrial ecosystems (Jordan, 1975). Extensive 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 Bristol, 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 and 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 emissions on terrestrial ecosystems near smelters 

and other industrial sites decrease 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; Spurgeon and Hopkin, 1996a; see also Section AX7.1.2). For example, 

the concentration of Pb in forest 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 structure and function, the extent of those effects 

is limited spatially. Elevated metal concentrations around smelters have been found to persist 

despite significant reductions 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 emissions 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 transport. 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 Brook Experimental Forest in New Hampshire, the 

amount of Pb in the forest floor was estimated 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 atmospheric Pb deposition from 1926 to 1987 to be 8.7 kg ha!1

, 

an amount that could account for nearly all of the increase in Pb in the forest floor during the 

period. The input of precipitation Pb to the 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 employing Pb budgets (Miller and Friedland, 1994; 

Watmough et al., 2004), and Pb isotopes (Bacon 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, represents a quantitatively significant fraction 

of labile Pb in temperate soils, especially in the upper, organic-rich horizons. 

Despite years of elevated atmospheric Pb inputs and elevated concentrations in soils, there 

is little evidence that sites affected primarily by long-range Pb transport have experienced 

significant effects on ecosystem structure or function. 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 Hubbard 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 solutions 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, whereas other elements (Pb, As, Se, Fe, Cr, Cu) 

are strongly sequestered, in the respective order noted. 
 AX7-99 

In ecosystems where Pb concentrations in soil solutions are low, toxicity levels for 

vegetation are not likely to be reached regardless of the soil Pb concentration. Furthermore, 

mycorrhizal infection of tree roots appears to reduce the translocation 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 infection of roots was not affected by Pb 

dose. Some, but not all, species of mycorrhizae 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 solution 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 vegetation. 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 Norway between 1982 and 1992, when atmospheric Pb 

deposition declined by approximately 70%. They also observed significantly lower Pb 

concentrations in tree tissues collected in northern Norway versus southern Norway, where 

atmospheric Pb deposition is greater. 

Even at subtoxic concentrations, Pb and other metals may influence species diversity in 

terrestrial ecosystems. However, little work has 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 authors hypothesized that Pb may inhibit phosphorous uptake by 

dominant species, allowing less abundant (but more Pb-tolerant) ones to succeed. 

AX7.1.4.3 Effects of Lead on Energy Flows and Biogeochemical Cycling 

In terrestrial ecosystems, energy flow is closely linked to the carbon cycle. The principal 

input of energy to terrestrial ecosystems is through photosynthesis, in which CO2 is converted to 
 AX7-100 

biomass carbon. Because of this link between photosynthesis and energy flow, any effect that Pb 

has on the structure and function of terrestrial ecosystems (as discussed 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 decomposition of litter and 

other detrital material by soil bacteria and fungi, and cascade through the various components of 

the detrital food web. Because the mobility of Pb in soils is closely tied to organic matter 

cycling, decomposition processes are 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 flows in terrestrial ecosystems. At some sites 

severely affected by metal pollution, death of vegetation can occur, 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; Galbraith et al., 1995). 

At Pb-affected sites that can retain a functioning forest stand, the rate of decomposition of 

litter may be reduced, resulting in greater accumulation of litter on the forest floor than in 

unpolluted stands. Numerous investigators have documented significant declines in litter 

decomposition rates (Cotrufo et al., 1995; Johnson 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-contaminated 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 from a smelter in Bristol, England had a litter 

layer mass 10 times greater than the mass in a similar stand 23 km from the smelter (Martin and 

Bullock, 1994). 

Reduced decomposition rates in polluted ecosystems 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 substrate-induced respiration in heavily polluted 

grassland soils near the U.S. Army’s Aberdeen Proving Ground in Maryland. This decline in 

carbon respiration was associated with 81% lower bacterial biomass and 93% lower fungal 
 AX7-101 

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,” 

qCO2) declined with increasing metal concentration, though this observation depended on the 

procedure for measuring microbial biomass (substrate-induced respiration versus fumigationextraction). The combined effect of lower microbial biomass per unit soil carbon and similar or 

lower qCO2 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 determining the effects of Pb and other metals 

on the microbial processing of litter. Forest soils in temperate zones accumulate organic matter 

at the soil surface to a greater degree than in grasslands. This organic-rich O horizon can support 

a large microbial biomass; but it is also an effective trap for Pb inputs, because of the association 

between Pb and soil organic matter. At highly contaminated 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 processes. Studies in the 1970s demonstrated that 

Pb and other metals inhibit the mineralization of nitrogen from soil organic matter and 

nitrification (Liang and Tabatabai, 1977, 1978), resulting in lower nitrogen availability to plants. 

More recent research has documented significant inhibitory 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 Tabatabai, 2000; Ekenler and Tabatabai, 2002). This 

suggests that the inhibitory effect of Pb and other metals is broad-based, and not specific to any 

particular metabolic pathway. In reducing environments, 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% decreases 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 earthworms 
 AX7-102 

in an oak woodland site 3 km from a Pb-Zn smelter in Bristol, England, compared 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 example, the abundance of the woodlice Trichoniscus pusillus

was 151 individuals per m2

 in the reference woodland, but none were found in the contaminated 

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

studied. At six sites within 1 km from the smelters, 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 invertebrate community. 

The effect of metal pollution on soil invertebrates may be a threshold-type response. In a 

study conducted in woodlands near two zinc smelters 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 control sites (Grelle et al., 2000). 

While Pb pollution affects the population and diversity 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 concentrations in earthworms were lower than soil Pb 

concentrations and much lower than litter Pb concentrations (Martin and Bullock, 1994). Litterdwelling mites had Pb concentrations that were 10% of the average litter 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 similar 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 their prey. Because of the absence of significant 

bioaccumulation in the soil food web, predator species will be affected by Pb pollution primarily 

through effects on the abundance of their 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 directly decreases the flow of energy into the 
 AX7-103 

ecosystems. More commonly, the accumulation of toxic levels of Pb or other metals in litter and 

soil decreases the rate of litter decomposition through decreases in microbial biomass and/or 

respiration. These reductions can subsequently affect higher trophic levels that depend on these 

organisms. It is important to note that sites that have exhibited significant disruption to energy 

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

adverse effects from SO2 from smelters or other metals-related activities. 

Lead Dynamics in Terrestrial Ecosystems 

Lead inputs to terrestrial ecosystems in the United 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 emissions controls. Unfortunately, there are few longterm data sets of precipitation inputs to terrestrial ecosystems. At the Hubbard Brook 

Experimental Forest, in New Hampshire, Pb input 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; Brännvall et al., 2001a,b). 

Reported concentrations of Pb in waters draining natural terrestrial ecosystems have 

always been low (Wang et al., 1995; Bacon and Bain, 1995; Johnson et al., 1995a; Vinogradoff 

et al., 2005), generally less than 1 ng L!1, even at moderately polluted sites (Laskowski et al., 

1995). Consequently, most terrestrial ecosystems in North America and Europe remain sinks for 

Pb despite reductions in atmospheric Pb deposition 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 retained in the terrestrial 

ecosystem in the later time period. Similar observations have been made in Europe, where the 

use of leaded gasoline has also declined in the 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 times 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 at 
 AX7-104 

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 reductions in emissions, this accumulation of Pb 

continues, though at markedly lower rates. The 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 Johnson, 1985; Turner et al., 1985). However, more 

recent literature suggests that Pb is transported more 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 sites 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 forest 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 significant 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 greatest at their sites in southern Vermont, 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 concentration 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 and Friedland, 1994; Johnson 

et al., 1995a; Johnson and Petras, 1998; Watmough and Hutchinson, 2004; Johnson et al., 2004). 

This is supported by evidence from Pb-isotope analyses. Gasoline-derived Pb has a 206Pb:207Pb 

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 accumulation 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 hardwood stand on Camel’s Hump Mountain in 

Vermont, as much as 65% of the pollutant Pb deposited to the stand had moved into mineral 
 AX7-105 

horizons by 2001 (Kaste et al., 2003). In a spruce-fir stand, containing a thicker organic forest 

floor layer, penetration of pollutant Pb into the mineral soil was much lower. 

This recent research has resulted in a reevaluation 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 published previously by Miller and Friedland (1994), 

who used a Pb budget approach. Extremely rapid 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 lower 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 Hutchinson (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 United States, pollutant Pb represents 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 observed 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, and 
 AX7-106 

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 terrestrial ecosystems exists at sites within about 

50 km of smelters and other Pb-emitting industrial sites. Assessing the risks specifically 

associated with Pb is difficult, because these sites also experience elevated concentrations of 

other metals and because of effects related to SO2 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-complete collapse of the detrital food web, creating a 

terrestrial ecosystem in which energy and nutrient flows are minimal. More commonly, stress in 

soil microorganisms and detritivores can cause reductions in the rate of decomposition of detrital 

organic matter. Although there is little evidence of significant bioaccumulation of Pb in natural 

terrestrial ecosystems, reductions in microbial and detritivorous populations can affect the 

success of their predators. Thus, at present, industrial point sources represent the greatest Pbrelated threat to the maintenance of sustainable, healthy, diverse, and high-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 consideration of laboratory toxicity studies and field evaluations. 

Perhaps the most straightforward approach for evaluating the effects of Pb is to consider extant 

criteria for Pb in aquatic ecosystems, i.e., water and sediment quality criteria. A key issue in 

developing Pb water and sediment criteria that are broadly applicable to a range of water bodies 

is properly accounting for Pb bioavailability and the range in species 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 concentrations are briefly addressed. It is beyond 

the scope of this section to review all methodologies in aquatic 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 ambient 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 low 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 239.1b, 

Standard Method 3111c

Graphite Furnace Atomic Absorption Spectroscopy 

(GFAAS) 

EPA SW-846 Method 7421 a, 

EPA Method 239.2b, 

Standard Method 3113c

Inductively Coupled Plasma 

(ICP) 

EPA SW-846 Method 6010B a, 

EPA Method 200.7b

, 

Standard Method 3120c

Inductively Coupled Plasma-Mass Spectrometry 

(ICP-MS) 

EPA SW-846 Method 6020 a, 

EPA Method 200.8b

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 aquatic 
 AX7-108 

ecosystems. Just as in the terrestrial environment, the speciation of Pb and other trace metals in 

natural freshwaters and seawater plays a crucial role in determining their reactivity, mobility, 

bioavailability, and toxicity. Many of the same speciation techniques employed for the 

speciation of Pb in terrestrial ecosystems (see Section AX7.1.1 and AX7.1.2) are applicable in 

aquatic ecosystems. 

There is now a better understanding of the potential effects of sampling, sample handling, 

and sample preparation on aqueous-phase metal speciation. 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 trace metal speciation techniques include: 

• Diffusion gradients in thin-film gels (DGT) 

• Gel integrated microelectrodes combined with voltammetric in situ profiling 

(GIME-VIP) 

• Stripping chronopotentiometry (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 compared in a study by Sigg et al. (2006) using 

freshwaters collected in Switzerland. They found that techniques 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 artifacts associated sampling and sample handling. 

AX7.2.1.2 Ambient Water Quality 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 laboratory. In general, both acute (short term) and chronic (long term) 

AWQC are developed. Depending on the species, the 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 

(LC50) or median effect concentration (EC50). For each species, the geometric mean of the 

acceptable LC50/EC50 data is calculated to determine the species mean acute value (SMAV). 
 AX7-109 

For each genera, the geometric mean of the relevant 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 percentile of the GMAVs, based on the four GMAVs surrounding the 

5th percentile) is determined. Because the FAV is based on LC50/EC50 values (which represent 

unacceptably high levels of effect), the FAV is divided by two to estimate a low-effect level. 

This value is then termed the acute criterion, or criterion 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 normalized 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 μg/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 partial life cycle toxicity 

studies are considered acceptable. 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-observed-effect concentration (NOEC)1

 and the lowest-observedeffect concentration (LOEC)2

. Although a chronic criterion could be calculated as the 5th 

percentile of genus mean chronic values (GMCVs), sufficient chronic toxicity 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 

LC50 to the chronic value, based on studies with the 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 for developing AWQC (Stephan et al., 1985) are 

now more than 20 years old and thus are not reflective of scientific advances in aquatic 

toxicology and risk assessment that have developed since the 1980s. For example, the 

toxicological importance of dietary metals has been increasingly recognized and approaches for 

incorporating dietary metals into regulatory criteria are being evaluated (Meyer et al., 2005). 

Other issues include consideration of certain sublethal endpoints that are currently not directly 

incorporated into AWQC development (e.g., endocrine toxicity, behavioral responses) and 

protection of threatened and endangered (T&E) species (U.S. Environmental Protection Agency, 

2003). In deriving appropriate and scientifically defensible air quality 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 summarize 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 did not result in statistically significant effects relative 

to the control. 

2

 The LOEC is the lowest concentration tested that resulted in statistically significant effects relative to 

the control. 
 AX7-111 

AX7.2.1.3 Ambient Water Quality Criteria: Bioavailability Issues 

In surface waters, the environmental fate of metal contaminants is mitigated through 

adsorption, complexation, chelation, and other processes that affect bioavailability. The toxicity 

of divalent cations tends to be highest in soft waters with low concentrations of dissolved organic 

matter and suspended particles. In an acidic environment (pH <4), the ionic 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 (Florence, 1977; Miller and Mackay, 1980). The portion of dissolved metal 

available for uptake or bioaccumulation is influenced 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 complexed (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 readily bioavailable and, thus, are less 

toxic to aquatic biota than their free-metal ion counterparts (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 and municipal and industrial effluent discharges 

include glycine, ammonia, oxalate, humic or fulvic acids, hydroxide, carbonate, bicarbonate, 

chloride, and hydrogen sulfide (Stumm and Morgan, 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 example, are 30, 65, and 136 μg/L at hardness levels of 

50, 100, and 200 mg/L (as CaCO3). Although it has been known for some time that other water 

quality parameters such as pH, dissolved organic carbon (DOC), and alkalinity affect the 

bioavailability of metals to aquatic biota, it was the relatively recent development of the biotic 

ligand model (BLM) that allows for AWQC to potentially consider all of these factors. Paquin 

et al. (2002) provided a thorough review of the factors influencing metal bioavailability and how 

research over the last few decades has culminated in the development of the BLM. 

By understanding the binding affinities of various natural ligands in surface waters and 

how the freshwater fish gill interacts with free cations in the water, one can predict how metals 

exert their toxic effects (Schwartz et al., 2004). Models developed prior to the BLM are the freeion activity model (FIAM) and the gill surface interaction model (GSIM). The FIAM accounts 
 AX7-112 

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 between metal species and other cations (Paquin 

et al., 2002). The GSIM is fundamentally similar to the FIAM in that it accounts for competition 

between metal ions and hardness cations at the physiological 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, therefore, considers that the external 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 within 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 ligand exceeds a threshold concentration (Di Toro 

et al., 2001; Paquin et al., 2002). Free metal cations “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 ionoregulatory mechanisms in the fish, 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 organisms, such as crustaceans, 

where the site of action is directly exposed to the aqueous environment (Di Toro et al., 2001). 

Although the BLM is currently being considered as a tool for 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 quality criteria for lead. For example, BLMs 

developed to-date have focused on acute mortality/immobilization endpoints for fish and 

invertebrates. Chronic exposures are typically of greatest regulatory concern, but chronic BLMs 

to date have received limited attention (De 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 biota and, in some cases, increased toxicity. Besser et al. (2005) 

observed that chronic (42-day) Pb toxicity to the amphipod Hyalella azteca was greater from a 

combined aqueous and dietary exposure than from a water-only exposure and the authors 

conclude that estimates of chronic toxicity thresholds 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 must be addressed in developing and applying a 
 AX7-113 

BLM in AWQC development is the type of dissolved organic matter used in model development 

relative to the types of dissolved organic matter at the site of interest. Richards et al. (2001) 

demonstrated that natural organic matter (NOM) of different types differentially influenced Pb 

and Cu accumulation by gills of rainbow trout. There are also other ligands not accounted for in 

the BLM that require more research. Bianchini and Bowles (2002) emphasized the importance 

of reduced sulfur as a metal ligand, limitations in scientific knowledge on reduced sulfur, and 

provided recommendations for studies 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 the 1985 AWQC were released and incorporation of the BLM is 

being evaluated. If an acute BLM is incorporated into the revised AWQC for lead, the chronic 

criteria would likely be estimated using an ACR (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 processes that reduce the concentration of free, 

bioavailable metal. Sediments function as a sink for Pb, as with most metals. Lead compounds 

such as Pb-carbonates, Pb-sulfates, and Pb-sulfides predominate in sediments (Prosi, 1989). 

Total Pb has a higher retention time and a higher percentage is retained in sediments compared to 

copper and zinc (Prosi, 1989). Lead is primarily 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 cation-exchangeable fractions and easily-reducible 

fractions are more mobile and more readily bioavailable to biota (Prosi, 1989). Most Pb 

transported in surface waters is in a particulate form, originating from the erosion of sediments in 

rivers or produced in the water column (Prosi, 1989). 

Sediment quality criteria have yet to be adopted by the EPA, but an equilibrium 

partitioning procedure has recently been published (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 metals 
 AX7-114 

partition in sediment between acid-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 metals. Note that this 

approach can be used to predict the lack of toxicity, but not the presence of toxicity. It is 

important to emphasize that metals must be evaluated as a mixture using this approach. 

If individual metals, or just two or three 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)/fOC), 

mortality can be more reliably predicted by accounting for both the site-specific organic carbon 

and AVS concentrations. When evaluating a metal mixture containing cadmium, copper, Pb, 

nickel, silver, and zinc, the following predictions can be made (U.S. Environmental Protection 

Agency, 2005c): 

• A sediment with (SEM / AVS)/fOC < 130 μmol/gOC should pose low risk of adverse 

biological effects due to these metals. 

• A sediment with 130 μmol/gOC < (SEM / AVS)/fOC < 3000 μmol/gOC may have adverse 

biological effects due to these metals. 

• In a sediment with (SEM / AVS)/fOC > 3000 μmol/gOC, adverse biological effects may 

be expected. 

A third approach is to measure pore water concentrations 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 has endorsed the AVS-SEM approach for 

developing sediment guidelines, other investigators do not necessarily agree, as the AVS-SEM 

approach may not be relevant to benthic organisms that ingest sediment particles. For example, 

Griscom et al. (2002) found that metals associated with either reduced or oxidized sediment 

particles can be assimilated by deposit and suspension 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 protecting some benthic organisms from acute 

toxicity associated with exposure to very high pore water concentrations of metals, they 

questioned whether the AVS-based approach is relevant to less contaminated sediments under 
 AX7-115 

more natural conditions with vertical stratification of oxygen concentrations and, hence, varying 

levels of AVS. The authors concluded that important 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 den Berg et al., 1998). Finally, Long et al. (1998) 

critically compared the ability of AVS-SEM and dry weight-normalized concentrations to predict 

toxicity of sediment-associated trace metals. Based on an analysis of 77 field-collected sediment 

samples, Long et al. (1998) found that dry weight-normalized concentrations were equally or 

slightly more accurate than AVS-SEM in predicting non-toxic and toxic results in laboratory 

bioassays. Long et al. suggests that a limitation of the AVS-SEM approach is that the criteria 

were not derived from field studies with mixture of toxic chemicals and thus may be less relevant 

to sites comprising complex chemical mixtures. Thus, although the AVS-SEM approach for 

developing sediment quality criteria 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 tests (Long et al., 1995; Ingersoll et al., 1996; MacDonald et al., 

2000). However, these guidelines are based on total metal concentrations in sediment and do not 

account for the bioavailability of metals between 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 Type 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 – median 
 AX7-116 

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 applied to AWQC, metal mixtures have been shown to be more 

toxic than individual metals (Spehar 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 additive for fathead minnows and nearly strictly 

additive for daphnids. One approach for considering the additive toxicity of Pb with other metals 

is to use the concept of toxic units (TUs). Toxic units for each component of a metal mixture are 

derived by dividing metal concentrations by their respective acute or chronic 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 same 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 model for predicting effects of metal mixtures based 

on the currently available toxicity data. However, it should also be emphasized that the TU 

approach is most appropriate at a screening level, because the true 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 sediments with other toxicants, such as organic 

contaminants. Effects-based sediment quality guidelines (SQGs) have been developed over the 

past 20 years to aid in the interpretation 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 calculating the mean of the quotients for the 

individual chemicals. Long et al. (2006) performed a critical review of this approach and found 

that it reasonably predicts the incidence and magnitude of toxicity in laboratory tests and the 

incidence of impairment to benthic communities increases incrementally with increasing 

mSQGQs. However, the authors pointed out some of the limitations of this approach, such as a 
 AX7-117 

lack of agreement on the level of mSQGQs, masking of an individual chemical’s effect due to 

data aggregation, lack of SQGs for all chemicals 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 remediation (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 from 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 biota. Background Pb concentrations are spatially 

variable depending on geological features and local characteristics 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 approach ignores the possible contribution of 

background metal levels to toxic effects, and background metal levels are regionally variable, 

precluding the approach from being easily transferable between sites. In terms of deriving 

environmental criteria for Pb, background levels should be considered on a site-specific basis if 

there is sufficient information that Pb concentrations are naturally elevated. As discussed 

previously, the use of radiogenic Pb isotopes is useful for source 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 erosional transport of soil particles (Baier and Healy, 1977; Dolske and 

Sievering, 1979; and Yang and Rose, 2005). A number of physical and chemical factors govern 

the fate and behavior of Pb in aquatic systems. The EPA summarized 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 (Pb2+) form and complexation 

with inorganic and organic ligands is dependent on pH (Lovering, 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 matter). The ratio of Pb in suspended solids to Pb in filtrate 
 AX7-118 

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 determined by temperature, pH, oxidation-reduction 

(redox) potential, organic content, grain size, and chemical form of Pb in the water and 

biological activities (Jenne and Luoma, 1977). Since the publication of the 1986 Lead AQCD 

(U.S. Environmental Protection Agency, 1986a), knowledge of the properties of Pb in aquatic 

ecosystems has expanded. This section will provide further detail on the chemical species and 

the environmental factors affecting speciation of Pb in the aquatic environment. In addition, 

quantitative distributions of Pb in water, sediment, 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 environment is controlled by many factors. The 

primary form of Pb in aquatic environments is divalent (Pb2+), while Pb4+ exists only under 

extreme oxidizing conditions (Rickard and Nriagu, 1978). Labile forms of Pb (e.g., Pb2+, 

PbOH+, PbCO3) are a significant portion of the Pb inputs to aquatic systems from atmospheric 

washout. Lead is typically present in acidic aquatic environments as PbSO4, PbCl4, ionic Pb, 

cationic forms of Pb-hydroxide, and ordinary Pb-hydroxide (Pb(OH)2). In alkaline waters, 

common species of Pb include anionic forms of Pb-carbonate (Pb(CO3)) 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 MINTEQA2 computer model 

(http://www.epa.gov/ceampubl/mmedia/minteq/) is one such equilibrium speciation model that 

can be used to calculate the equilibrium composition of dilute aqueous 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 aquatic biota and was previously 

described in Section AX7.2.1.3. Further detail on speciation models 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 ecosystems is dependent upon pH and the 

available complexing ligands. Solubility varies according to pH, temperature, and water 

hardness (Weber, 1993). Lead rapidly loses solubility above pH 6.5 (Rickard and Nriagu, 1978) 

and as water hardness increases. In freshwaters, Pb typically forms strong complexes with 

inorganic OH!and CO3

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 Pb2+ and less abundant 

inorganic forms include Pb(HCO)3, Pb(SO4)2

2!

, PbCl, PbCO3, and Pb2(OH)2CO3

(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., nitrilotriacetonitrile and ethylenediaminetetraacetic acid 

[EDTA]) (U.S. Environmental Protection Agency, 1986b). The presence of organic complexes 

has been shown to increase the rate of solution 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 decomposed. 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 interactions between metals and humic acids. 

Marine Water

In marine systems, an increase in salinity increases complexing with chloride and 

carbonate ions and reduces the amount of free Pb2+. In seawaters and estuaries at low pH, Pb is 

primarily bound to chlorides (PbCl, PbCl2, PbCl3

!

, PbCl4

2!

) and may also form inorganic 

Pb(HCO)3, Pb(SO4)2

2!

, or PbCO3. 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 PbCl3

!

 > PbCO3 > PbCl2 > PbCl+ > and Pb(OH)+ (Fernando, 1995). 

Lead in freshwater and seawater systems is highly complexed with carbonate ligands suggesting 

that Pb is likely to be highly available for sorption to suspended materials (Long and Angino, 

1977). 

Current information suggests that inorganic Pb is the dominant form in seawater; 

however, it has been shown that organically bound Pb complexes make up a large portion of the 

total Pb (Capodaglio et al., 1990). 

Sorption 

Sorption processes (i.e., partitioning of dissolved 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). Sorption of Pb results in the enrichment of 
 AX7-121 

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 adsorption to aquatic sediments is correlated with pollution in sites 

containing high levels of anthropogenic organic content, even under acidic conditions (Tada and 

Suzuki, 1982; Brook and Moore, 1988; Davis and Galloway, 1993; Botelho et al., 1994; Davis 

et al., 1996). Particulate-bound forms are more often linked to urban runoff and mining effluents 

(Eisler, 2000). 

Solid Pb complexes form when Pb precipitates or adsorbs to suspended particulate matter 

and sediments. Inorganic Pb adsorption to suspended organic matter or sediments is dependent 

on parameters such as, pH, salinity, water hardness, 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., clays). Adsorption decreases with water hardness 

(Syracuse Research Corporation, 1999). At higher pH, Pb precipitates as Pb(OH)+

 and PbHCO3

+

into bed sediments (Weber, 1993). Conversely, at low pH, Pb is negatively sorbed (repelled 

from the adsorbent surface) (U.S. Environmental Protection Agency, 1979; Gao et al., 2003). 

In addition, Pb may be remobilized from sediment with a decrease in metal concentration in the 
 AX7-122 

solution phase, complexation with chelating agents (e.g., EDTA), and changing redox conditions 

(Gao et al., 2003). Changes in water chemistry (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 remobilization and reintroduction into the aqueous 

environment compartment and its subsequent release into the atmosphere (Syracuse Research 

Corporation., 1999). However, methylation is not a significant environmental pathway 

controlling the fate of Pb in the aquatic environment. The microbial methylation of Pb in aquatic 

systems has been demonstrated experimentally, but evidence for natural 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 bacteria known to alkylate metals and did not find 

evidence of Pb methylation under any test condition. Tetramethyl-Pb may be formed by the 

methylation of Pb-nitrate or Pb-chloride in sediments (Bodek et al., 1988). However, 

tetramethyl-Pb is unstable and may degrade in aerobic environments after being released from 

sediments (U.S. Environmental Protection Agency, 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 methylation of Pb compounds and suggested that 

chemical methylation of Pb is the dominant process 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 Protection Agency, 1986b) did not describe 

the distribution and concentration of Pb throughout aquatic ecosystems of the United States. 

Consequently, an analysis of readily available data on Pb concentrations was conducted to 

determine the distribution of Pb in the aquatic environment. Data from the United States 

Geological Survey (USGS) National Water-Quality 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 river basins and aquifers throughout the country, 
 AX7-123 

and it has been used by the EPA for describing national environmental concentrations for use in 

developing AWQC. The authors recognize that the NAWQA program encountered analytical 

challenges with the chemical analysis of Pb in surface waters. The analytical methods available 

during the NAWQA program were not as sensitive as methods currently applied today. 

Therefore, analytical detection limits are elevated 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 supplemented 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 during long-term, cyclical investigations wherein study units 

undergo intensive sampling for 3 to 4 years, followed by low-intensity monitoring and 

assessment of trends every 10 years. The NAWQA 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 available through 30 September 2003. The NAWQA program study 

units were selected to represent a wide variety of environmental conditions and contaminant 

sources; therefore, agricultural, urban, and natural areas were all included. Attention was also 

given to selecting sites covering a wide variety of hydrologic and ecological resources. 

NAWQA sampling protocols are designed to promote data consistency within and among 

study units while minimizing local-scale spatial variability. 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 hydrologic 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 along a 100-m stream reach, 

upstream of the location of the water-column sampling. Five to ten depositional zones at various 

depths, covering left bank, right bank, and center channel, are sampled to ensure a robust 

representation of each site. Fine-grained 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 decision trees that help guide selection from that list to 

accommodate local variability. 
 AX7-124 

The NAWQA dataset was chosen over other readily available national databases (i.e. the 

USEPA-maintained database for the STOrage and RETrieval [STORET] of chemical, physical, 

and biological data), because the study design and 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 protocol has lead to the STORET data being highly qualified 

and offered with only a mild level 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 the available Pb data in STORET predate the use 

of clean techniques for Pb quantification. 

The authors recognize the existence of several local and regional datasets that may be of 

quality equal to NAWQA; however, due to the national scope of this assessment, these datasets 

were not included in the following statistical analyses. However, because the NAWQA database 

does not cover lakes and the marine/estuarine environment, 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 entire United States (all states) from the 

NAWQA website (http://water.usgs.gov/nawqa/index.html): site information, dissolved Pb 

concentration in surface water (μg/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 the ambient group (the combined contribution of 

natural and anthropogenic sources), whereas the natural group comprised “forest,” “rangeland,” 

or “reference” samples only4

. These groups follow those defined 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 that while Pb samples collected from sites classified under these three land use 

categories will most closely reflect natural background concentrations, 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 natural/ambient classification, tissue samples were 

further divided into “whole organism” and “liver” groups. 

All data were compiled in spreadsheets wherein non-detect values were converted to onehalf the detection limit and the total number of samples, percentage of non-detect values (percent 

censorship), minimum, maximum, median, mean, 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 the mean to a greater extent than inclusion of 

non-detects as 1⁄2 of the detection limit (Newman 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 analyses, and calculate statistics. 
 AX7-126 

Lead Distributions Generated 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 limited. 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 samples except in the Midwest (Ohio, Illinois, Iowa, and 

Michigan), where natural sites were not present (Figure AX7-2.2.3). This exception may be 

because these areas are dominated by agricultural 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 5.44 
 AX7-127 

Table AX7-2.2.3. Summary Statistics of Ambient 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 samples 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 limit5 (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 falling at 0.50 μg/L (one-half 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 percentiles 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 natural 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 detection 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 concentrations 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. 
 AX7-128 

Figure AX7-2.2.3. Spatial distribution of natural and ambient surface water/sediment sites (Surface wate

ambient N = 3445; Sediment: natural N = 258, ambient N = 1466). 
 AX7-129 

Figure AX7-2.2.4. Spatial distribution of natural and ambient liver tissue sample sites (Natural N = 83, 
 AX7-130 

Figure AX7-2.2.5. Spatial distribution of natural and ambient whole organism tissue sample sites (Natu

Ambient N = 332). 
 AX7-131 

Figure AX7-2.2.6. Frequency distribution of ambient 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 detectable Pb concentrations above 

and below the 99th percentile. The 99th percentile (versus the 95th percentile) was chosen in 

this instance to represent extreme conditions given the small window of variability in the dataset. 

By convention, the 95th percentile was 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 measured 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 limit (7/1466 ambient ND, 3/258 natural ND; Table AX7-2.2.3). 

As expected, the mean ambient Pb concentration was higher than the mean natural Pb 
 AX7-132 Figure AX7-2.2.7. Spatial distribution of dissolved lead in surface water (N = 3445). 

`
 AX7-133

concentration (120.11 and 109.07 μg/g, respectively). 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 percentile was higher than the natural 95th percentile 

(200.00 and 161.50 μg/g, respectively). While the natural and ambient surface water Pb 

distributions differed only at the extremes, the natural sediment Pb percentiles were consistently 

lower than the ambient percentiles throughout the distributions (Figure AX7-2.2.8). Unlike the 

surface water dataset, because the sediment dataset was not heavily censored, assessing national 

trends in sediment Pb concentrations was possible. The data were mapped and categorized into 

the four quartiles of the frequency distribution (Figure AX7-2.2.9). The following observations 

were made: 

• Sediment Pb concentrations generally increased from west to east (the majority of 

sites along East Coast had Pb concentrations in the fourth quartile of the sediment Pb 

concentration frequency distribution). 

• Several “hot spots” of concentrated sites with elevated sediment Pb concentrations 

were apparent in various western states. 

• Sediment Pb concentrations were generally 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 concentrations, the highest measured sediment Pb 

concentrations were found in Idaho, Utah, and Colorado. Not surprisingly, of the top 10 

sediment Pb concentrations recorded, 7 were measured at sites classified as mining land use. 

Tissue 

As was true for the surface water data, there were a high number of tissue samples below 

the detection limit (47/93 natural 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 organism samples than liver samples, and for 

ambient sites than natural sites. As expected, for whole organism samples, the 95th percentile Pb 

concentration measured at ambient sites was higher than that measured at natural sites (3.24 and 

2.50 μg/g, respectively); however, Pb liver concentration 95th percentiles for ambient and 

natural samples were very similar, with the natural 95th percentile actually higher than the 
 AX7-134

Figure AX7-2.2.8. Frequency distribution of ambient and natural levels of bulk sediment 

<63 μm total Pb (μg/g). 

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

and mean Pb liver concentrations of ambient samples (0.15 and 0.36 μg/g, respectively) were 

higher than the median and mean Pb liver concentrations 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: median = 0.59, mean = 1.03; natural: median = 0.35, mean = 

0.95 μg/g). In addition, the frequency distributions of the liver and whole organism Pb 

concentrations followed the same trends, with the natural percentiles consistently lower than the 

ambient percentiles throughout the distributions (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 Contaminant Biomonitoring Program (NCBP) (Schmitt and 

Brumbaugh, 1990). As part of this program, 321 composite 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 was 
 AX7-135 Figure AX7-2.2.9. Spatial distribution of total lead 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 compare 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 assessment of spatial trends in Pb tissue concentrations. Instead, 

areas with high Pb tissue concentrations were identified by classifying the data above and below 

the 95th percentile. Similar to surface water and sediments, tissue concentrations were found to 

be elevated in Washington, Idaho, Utah, Colorado, Arkansas, and Missouri; however, several of 

the highest measured Pb concentrations were also found in study units in the southwestern and 
 AX7-137 

Figure AX7-2.2.10. Frequency distribution of ambient and natural levels of lead in liver 

tissue (μg/g dry weight). 

Figure AX7-2.2.11. Frequency distribution of ambient 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 NCBP1 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 weight values were multiplied by a factor of five. 

southeastern states (Figure AX7-2.2.12 and Figure 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 cover lakes or the sea where atmospheric 

deposition of Pb is highly likely, the primary literature 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 water 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 surface water dissolved Pb concentrations measured 

in the Great Lakes (Superior, Erie, and Ontario) between 1991 and 1993 were 3.2, 6.0, and 

9.9 ng/L, respectively. Lead concentrations ranged from 3.2 – 11 ng/L across all three lakes. 

Similarly, 101 surface water total 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 concentrations measured at these locations would 

likely have been even lower than the total Pb concentrations reported. 
 AX7-139 Figure AX7-2.2.12. Spatial distribution of lead in liver tissues (N = 559). 
 AX7-140 Figure AX7-2.2.13. Spatial distribution of lead in whole organism tissues (N = 332). 
 AX7-141 

In open waters of the North Atlantic the decline of Pb concentrations has been associated 

with the phasing out of leaded gasoline in North 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 Swamp in southern Georgia/northern Florida 

(Jackson et al., 2004). Based on sediment cores from the Okefenokee Swamp, Pb concentrations 

were approximately 0.5 mg/kg prior to industrial development, reached a maximum of 

approximately 31 mg/kg from about 1935 to 1965, and 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 sediment cores from 35 reservoirs and lakes in 

urban and reference settings were analyzed by Mahler et al. (2006) to determine the effects of 

three decades of legislation, regulation, and changing 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 were for Pb and Cr: For Pb, 83% of the lakes had 

decreasing trends and 6% had increasing trends; for Cr, 54% of the lakes had decreasing trends 

and none had increasing trends. Mass accumulation rates of metals in cores, adjusted for 

background concentrations, decreased from the 1970s to the 1990s, where median changes 

ranged from 246% (Pb) to 23% (Hg and Zn). The 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. However, Mahler et al. (2006) found 

that anthropogenic mass accumulation rates in dense 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 urban fluvial source signals can overwhelm those from regional 

atmospheric sources. In estuarine systems, however, 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 attributed 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 balance 
 AX7-142 

calculations indicate that only a small fraction of leaded gasoline fallout from the late 1980s has 

been washed out of the San Joaquin and Sacramento rivers’ drainage basin by 1995 and 

consequently freshwater inputs remain a Pb source to the bay (Steding et al., 2000). The authors 

suggest that the continuous source of Pb from the river systems draining into the bay, coupled 

with benthic remobilization of Pb, indicates that historic gasoline deposits may remain in the 

combined riparian/estuarine system for decades. 

AX7.2.2.3 Tracing the Fate and Transport of Lead in Aquatic Ecosystems 

The following section presents a generalized framework for the fate and transport of Pb in 

aquatic systems (Figure AX7-2.2.14). The primary source of Pb in natural systems is 

atmospheric deposition (Rickard and Nriagu, 1978; U.S. Environmental Protection Agency, 

1986a). Estimated median global atmospheric emission for anthropogenic and natural sources 

are 332 Η 106

 kg/year and 12 Η 106 kg/year, respectively (summarized by Giusti et al., 1993). 

Inorganic and metallic Pb compounds are nonvolatile and will partition to airborne particulates 

or water vapors (Syracuse Research Corporation., 1999). Dispersion and deposition of Pb is 

dependent on the particle size (U.S. Environmental Protection Agency, 1986a; Syracuse 

Research Corporation., 1999). More soluble forms 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, erosion, or direct point source inputs 

(e.g., Pb shot or accidental spills). Once in the 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 particulate 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 assimilated by aquatic organisms, or be resuspended in the 

water column. The uptake, accumulation, and toxicity of Pb in aquatic organisms from water 

and sediments are influenced by various environmental factors (e.g., pH, organic matter, 

temperature, hardness, bioavailability). These factors are further described in Section 

AX7.2.3.4). The remainder of this section discusses 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 concentrations in various aquatic 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. Sediment concentration profiles are 

typically coupled with Pb isotopic analysis. The isotope fingerprinting method utilizes 

measurements of the abundance of common Pb isotopes (i.e., 204Pb, 206Pb, 207Pb, 208Pb) to 

distinguish between natural Pb over geologic time and potential anthropogenic sources. Details 

of this method were described in Section AX7.1.2. The concentration of isotope 204Pb has 

remained constant throughout time, while the other isotope species can be linked to various 

anthropogenic Pb sources. Typically, the ratios or signatures of isotopes (e.g., 207Pb:206Pb) 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 vary with depth. For example, Chow et al. 

(1973) examined sediment Pb profiles in southern California. Lead concentrations were 

increased in the shallower sediment depths and comparatively decreased at greater depths. These 

changes in sediment vertical concentration were attributed to higher anthropogenic Pb fluxes 
 AX7-144 

from municipal sewage, storm runoff, and atmospheric deposition. Similar experiments 

conducted throughout the United States have also suggested an increase in Pb concentrations in 

the upper sediment layer concomitant with increases in anthropogenic inputs (Bloom and 

Crecelius, 1987; Case et al., 1989; Ritson 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 Bay, CA to a specific slag deposit. Several 

investigators have examined isotopic tracers to determine potential regional sources of Pb in 

eastern North America and the Great Lakes (Flegal et al., 1989b; Graney et al., 1995; Blais, 

1996). Water samples from Lake Erie and Lake Ontario were collected and analyzed. Lead 

isotope ratios (206Pb:207Pb) from the lakes were compared to known ratios for Pb aerosols derived 

from industrial sources in Canada and the United States and found to correlate 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 206Pb:207Pb ratios in sediment cores collected 

from a 300 km transect in Canadian Shield headwater 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-derived 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). Examination of Pb isotopes in sediment and 

suspended sediment in the St. Lawrence River were used to identify potential 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 gasoline consumption from 1930 to 1980. 

The historic record of atmospheric Pb pollution 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 Sweden (Bindler et al., 1999). In this study, soil, 

sediment, and tree rings were sampled for Pb concentrations and isotopic analyses were 
 AX7-145 

conducted on the soil samples. From this record, historic Pb concentrations and Pb accumulation 

rates were estimated. Present day concentrations in the forest soils ranged from 40 to 100 mg/kg, 

while a natural background concentration was estimated at <1 mg/kg. The authors were able to 

model Pb concentrations for the past 6,000 years and also to project Pb concentrations for the 

next 400 years, given an assumed atmospheric 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 assessed through retrospective measurements. 

Squire et al. (2002) used a time-series 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 changes 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 identified a 40% decline of Pb in the southern portion of the bay but 

found no change in the northern reach. 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, predominantly originating from 

atmospheric deposition or point source contribution. The fate and behavior of Pb in aquatic 

systems is regulated by physical and chemical factors such as pH, salinity, sediment sorption, 

transformation, and uptake by aquatic biota. In the United States, Pb concentrations in surface 

waters, sediments, and fish tissues range from 0.04 to 30 μg/L, 0.5 to 12,000 mg/kg, and 0.08 to 

23 mg/kg, respectively. Atmospheric sources are generally decreasing, as the United States has 

removed Pb from gasoline and other products. However, elevated Pb concentrations remain at 

sites associated with mining wastes or wastewater effluents. Since the 1986 Lead AQCD, much 

has been learned about the processes affecting Pb fate and transport. Detailed analyses are 

currently available (i.e., Pb isotope dating) to allow for constructing the history of Pb 

accumulation and identifying specific Pb contaminant sources. Continued source control along 

with examination of the physical and chemical properties will further allow 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 responses of aquatic biota to Pb exposure are 

highlighted in this section. A summary of the conclusions 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 conclusions and recommendations contained in 

the EPA staff review of that document (U.S. Environmental Protection Agency, 1990) are also 

provided. In addition, this section summarizes research subsequent to the 1986 Lead AQCD on 

Pb uptake into aquatic biota, effects of Pb speciation on uptake, resistance mechanisms to Pb 

toxicity, physiological effects of Pb, factors that affect responses to Pb, and factors associated 

with global climate change. Areas of research that are not addressed here include literature 

related to exposure to Pb shot or pellets and studies that examine 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 nonessential and non-beneficial and is toxic to living organisms in all 

of its forms (Eisler, 2000). Lead can bioaccumulate in the tissues of aquatic organisms through 

ingestion of food and water and adsorption from 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 solubility of Pb in water, dietary Pb (i.e., lead 

adsorbed to sediment, particulate matter, and food) may contribute substantially to exposure and 

toxicity in aquatic biota (Besser et al., 2005). Besser et al. (2004) exposed the amphipod 

Hyalella azteca to concentrations of Pb to evaluate the influence of waterborne and dietary Pb 

exposure on acute and chronic toxicity. The authors found that acute toxicity was unaffected by 

dietary exposure but that dietary Pb exposure did contribute to chronic toxic effects (i.e., 

survival, growth, reproduction) in H. azteca. Field studies in areas 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 demonstrated the effects of dietary metals on rainbow trout. 

However, there has been a debate on the importance 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 differences in the availability of 

Pb from the dietary sources used in laboratory studies, differences in speciation, and/or 
 AX7-147 

nutritional characteristics of the Pb dosed diets. In many field and laboratory studies, dietary 

exposure is rarily considered, but food provided to biota in these studies adsorb metals from 

water. Therefore, both dietary and waterborne exposure are occurring and both may be 

considered to play roles in eliciting the measured effects. 

Lead concentrations in the tissues of aquatic organisms are generally higher in algae and 

benthic organisms and lower in higher trophic-level consumers (Eisler, 2000). Thus, trophic 

transfer of Pb through food chains is not expected (Eisler, 2000). Metals are not metabolized; 

therefore, they are good integrative indicators of exposure in aquatic biota (Luoma and Rainbow, 

2005). Metal uptake is complex, being influenced by geochemistry, route of exposure (diet and 

adsorption), depuration, and growth (Luoma and Rainbow, 2005). This section discusses the 

factors affecting uptake of Pb by aquatic biota and the state of current research in this area. 

As described in Section AX7.2.2.1, the solubility of Pb in water varies with pH, 

temperature, and ion concentration (water hardness) (Weber, 1993). Lead becomes soluble and 

bioavailable under conditions of low pH, organic 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 precipitates out as Pb(OH)+

and PbHCO3

+

 into bed sediments. However, at reduced pH levels or ionic concentrations, 

sediment Pb can remobilize and potentially become bioavailable to aquatic organisms (Weber, 

1993). 

The most bioavailable inorganic form of Pb is divalent Pb (Pb2+), 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 movement 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 accumulation. 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 kept at pH 7.5. However, Albers and Camardese 

(1993a,b) examined the effects of pH on Pb uptake in aquatic plants 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 by 
 AX7-148 

aquatic plants and insects and on the concentration 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 metal form in the tissues of prey items (Luoma 

and Rainbow, 2005). Lead is typically present in acidic aquatic environments as PbSO4, PbCl4, 

ionic Pb, cationic forms of Pb-hydroxide, and ordinary hydroxide Pb(OH)2. In alkaline waters, 

common species of Pb include anionic forms of Pb-carbonate (Pb(CO3)) and Pb(OH)2. Labile 

forms of Pb (e.g., Pb2+, PbOH+

, PbCO3) are a significant portion of the Pb inputs to aquatic 

systems from atmospheric washout. Particulate-bound forms are more often linked to urban 

runoff and mining effluents (Eisler, 2000). Little research has been done to link the complex 

concepts of chemical speciation and bioavailability in natural systems (Vink, 2002). The 

relationship between the geochemistry of the underlying sediment and the impact of temporal 

changes (e.g., seasonal temperatures) to metal speciation are particularly not well studied (Vink, 

2002; Hassler et al., 2004). 

Generally speaking, aquatic organisms exhibit three Pb accumulation strategies: 

(1) accumulation of significant Pb concentrations with a low rate of loss, (2) excretion of Pb 

roughly in balance with availability of metal in the environment, and (3) weak net accumulation 

due to very low metal uptake rate and no significant excretion (Rainbow, 1996). Species that 

accumulate nonessential metals such as Pb and that have low rates of loss must partition it 

internally in such a way that it is sparingly available metabolically. Otherwise, it may cause 

adverse toxicological effects (Rainbow, 1996). Aquatic organisms that exhibit this type of 

physiological response have been recommended for use both as environmental indicators of 

heavy metal pollution (Borgmann et al., 1993; Castro 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 and Zolotukhina, 1989; Simòes Gonçalves et al., 

1991; Carter and Porter, 1997). 

Uptake experiments with aquatic plants and invertebrates (e.g., macrophytes, 

chironomids, crayfish) have shown steady increases 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 reduction in body Pb concentrations, as Pb incorporated into the 

crayfish shell is eliminated (Knowlton et al., 1983). Vázquez et al. (1999) reported on the uptake 
 AX7-149 

of Pb from solution to the extracellular and intracellular compartments of 3 species of aquatic 

bryophytes. Relative to the 6 metals tested, Pb was 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 metals were defined as metals 

introduced into the cell through a metabolically 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 abalones had generally lower whole 

body concentrations of heavy metals than did younger, 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 reach 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) suggested that mature physiological systems 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 rate of loss as described above. There are 

insufficient data available to determine whether 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 individual the more the metal content 

is diluted by body tissue (Rainbow, 1996). 

Once Pb is absorbed, it may sequester into varying parts of the organism. Calcium 

appears to have an important influence on Pb transfer. 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 natural systems, a model incorporating some of 

the factors mentioned above is desirable. The EPA’s Environmental Research Laboratory in 

Duluth, Minnesota developed a thermodynamic equilibrium 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 specifically 
 AX7-150 

designed to model uptake, MINTEQ provides an indication of what forms of the metal are likely 

to be encountered by aquatic organisms by estimating the formation of metal ions, complexation 

of metals, and the general bioavailability of metals from environmental parameters. More 

recently, a mechanistic model centered on biodynamics 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 potential 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 metal bioavailability and its link to effects 

(Hassler et al., 2004). These include the free ion activity model (FIAM) and the biotic ligand 

model (BLM). Specific information on these models, including their limitations, is provided in 

Section AX7.2.1.3. These models are useful for 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 various aquatic plants in Table 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). Duckweed that was either previously exposed or 

not exposed to Pb was exposed to a single dose of Pb-nitrate at 5000 μg/L for 21 days. 

Duckweed that was previously exposed to Pb removed 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 invertebrates in Table AX7-2.3.2. BCFs for 

freshwater snails were 738 for a 28-day exposure (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 (MacLean et al., 1996), and BCFs of 499 and 1120 for 

the caddis fly and stonefly, respectively, in 28-day exposures (Spehar et al., 1978). In a 28-day 

exposure, midge larvae were reported with a BCF of 3670 (Timmermans 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 Factors 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; Atchison et al., 1977). Although no BCFs have been reported for 

amphibians, Pb-nitrate was reported to accumulate mainly in the ventral skin and in the kidneys 

of frogs (Vogiatzis and Loumbourdis, 1999). 

Bioconcentration factors and bioaccumulation factors (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 BCFs and BAFs in organisms may not accurately 

reflect how metals are treated within the organisms (e.g., partitioning to specific organelles, 
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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 processes 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 frequently involves the sequestration of the metal, 

rendering it metabolically inactive. Recent research into heavy metal detoxification in aquatic 

biota has focused on several physiological and biochemical mechanisms for detoxifying Pb. 

This section examines these mechanisms and the ability of plants, protists, invertebrates, and fish 

to mitigate Pb toxicity. 

Plants and Protists 

Deng et al. (2004) studied the uptake and translocation 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 amount 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 varied within a narrow range. 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 and Davis (1977). Some 

of the wetland plants examined by Deng et al. (2004) also accumulated 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-containing intracellular metal-binding polypeptides that are 

produced by plants and protists in response to excessive uptake of heavy metals (Zenk, 1996). 

Phytochelatins are synthesized by the enzyme phytochelatin synthase that is activated by the 

presence of metal ions and uses glutathione as a substrate. When phytochelatins are synthesized 
 AX7-153 

in sufficient amounts to chelate the metal ion, the enzyme is deactivated (Morelli and Scarano, 

2001). 

Morelli and Scarano (2001) studied phytochelatin 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-phytochelatin complex content was lost. Their findings 

support a hypothesis of vacuolarization proposed for higher plants (Zenk, 1996), in which metalphytochelatin complexes are actively 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 and Scarano (2001) proposed concomitant occurrence 

of phytochelatin synthesis and release during metal exposure, as a coincident detoxification 

mechanism in P. tricornutum. 

Aquatic Invertebrates 

Like plants and protists, aquatic animals detoxify Pb by preventing it from being 

metabolically available, though their mechanisms for doing so vary. Invertebrates use 

lysosomal-vacuolar systems to sequester and process Pb within glandular cells (Giamberini and 

Pihan, 1996). They also accumulate Pb as deposits 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 radiotracers to evaluate the transfer of Pb into the food pathway 

of the starfish Asterias rubens as well as its distribution 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 sequestered 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 suggest that Pb migrates 

within the body wall from the organic matrix to the calcified skeleton. From there, the metal is 

either absorbed directly or adsorbed on newly produced ossicles (small calcareous skeletal 

structures), where it is efficiently retained as mineral deposition and is not metabolically active 

(Boisson et al., 2002). 
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AbdAllah and Moustafa (2002) studied the Pb storage capability of organs in the marine 

snail Nerita saxtilis. Enlarged electron-dense vesicles and many granules were observed in 

digestive cells of these snails and are suggested to be the site of storage of detoxified metals. 

N. saxtilis were found to be capable of concentrating Pb up to 50 times that of surrounding 

marine water without exhibiting signs of histopathologic changes. This ability has been 

attributed to chelation with various biochemical compounds, such as thionine (forming 

metallothionine) (Rainbow, 1996), or complexation with carbonate, forming lipofuchsin 

(AbdAllah and Moustafa, 2002). Granules observed in lysosomal residual bodies were presumed 

to be the result of Pb accumulation. The presence of large vacuoles and residual bodies were 

indicative of the fragmentation phase of digestion, suggesting that Pb was also processed 

chemically in the digestive cells. 

The podocyte cells of the pericardial gland of bivalves 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-vacuolar 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 detoxification mechanism of the tiger prawn Penaeus 

monodon. Vogt and Quinitio (1994) found that Pb granules tended to accumulate in the 

epithelial cells of the antennal gland (the organ of excretion) of juveniles exposed for 5 and 

10 days to waterborne Pb. The metal is deposited in vacuoles belonging to the lysosomal 

system. Continued deposition leads to the formation of electron-dense granules. Mature 

granules are released from the cells by apocrine secretion into 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 efficient 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; Anderson 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, showed 
 AX7-155 

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 hepatopancreas is the organ of metal 

storage and detoxification, although the molecular mechanisms of metal balance in crayfish have 

yet to be extensively investigated (Anderson et al., 1997). 

Fish

Most fish use mucus as a first line of defense against heavy metals (Coello and Khan, 

1996). In fish, some epithelia are covered with extracellular mucus secreted from specialized 

cells. Mucus contains glycoproteins, and composition varies among species. Mucosal 

glycoproteins chelate Pb, and settle, removing the metal from the water column. Fish may 

secrete large amounts of mucus when they come into contact with potential chemical and 

biochemical threats. Coello and Khan (1996) investigated the role of externally added fish 

mucus and scales in accumulating Pb from water, and the relationship of these 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 various parts of the body. Metallic Pb stimulated 

filamentous secretion, mostly from the ventrolateral 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 LT50 (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 LT50 than mucous. Fish scales can 

accumulate high concentrations of metals, including Pb, through chelation with keratin. Scales 

were shown to buffer the pH of Pb-nitrate in solution and remove Pb from water after which they 

settled out of the water column. Addition of scales to test water made all species (green sunfish, 

goldfish, and largemouth bass) more tolerant of Pb. 
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Summary of Detoxifiction Processes

Mechanisms of detoxification vary among aquatic biota and include processes such as 

translocation, excretion, chelation, adsorption, vacuolar storage, and deposition. Lead 

detoxification has not been studied extensively in aquatic organisms, but existing results indicate 

the following: 

• Protists and plants produce intracellular polypeptides that form 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 (Knowlton et al., 1983; Deng et al., 

2004). 

• Some starfish (asteroids) sequester the metal via mineral deposition into the exoskeleton 

(Boisson et al., 2002). 

• Species of mollusc employ lysosomal-vacuolar systems that store and chemically process Pb 

in the cells of their digestive and pericardial glands (Giamberini and Pihan, 1996; AbdAllah 

and Moustafa, 2002). 

• Decapods can efficiently excrete Pb (Vogt and 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 potentially reduce visceral 

exposure. 

Avoidance Response 

Avoidance is the evasion of a perceived threat. Recent research into heavy metal 

avoidance in aquatic organisms has looked at dose-response 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 examines preference-avoidance responses of 

invertebrates and fish to Pb and some other environmental gradients. 

Aquatic Invertebrates

Only one study was identified on avoidance response 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 metals 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) and 
 AX7-157 

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 (Lefcort et al., 2004). These results 

are consistent with those from similar studies. Control snails from reference sites, and first and 

second-generation snails from contaminated sites were capable of detecting and avoiding heavy 

metals, although the first generation was better than the second generation, and the second was 

better than the controls at doing so. This suggests 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) studied the preference-avoidance 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 experiment, tadpoles were exposed to an influx of 1000 μg Pb/L at five 

different infusion rates (i.e., volumes per unit time into the test system). Experiments were 

videotaped and location data from the tank were used to assess response. No significant 

differences were seen in preference-avoidance responses to Pb in either nonexposed or 

previously exposed animals. In a similar subsequent study, 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 physiological stress (Steele et al., 1991). 

The olfactory system in fish is involved in their forming avoidance response to heavy 

metals (Brown et al., 1982; Svecevičius, 1991). It is generally thought that behavioral avoidance 

of contaminants may be a cause of reduced fish populations in some water bodies, because of 

disturbances in migration and distribution patterns (Svecevičius, 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) studied metal mixture avoidance response in brown trout 

(Salmo trutta), as well as the added effects of acidification. A 1-fold (1Η) mixture contained 
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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 spent in test water, trip time to test water, and number of trips. 

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

0.1Η mixture. Reduced avoidance was observed at higher concentrations (4Η and 10Η). The 

authors proposed that the reduced avoidance response 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 study does not conclusively indicate which of the 

metals in the mixture may be causing the avoidance response. However, 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 1Η

metal mixture treatment, brown trout made fewer trips 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 the preference-avoidance responses of lake 

whitefish (Coregonus clupeaformis) to overlapping gradients of light 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 lacked 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 concentrations at or above 32,000 μg/L. 

Summary of Avoidance Response

In summary, of those aquatic organisms studied, 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 concentrations. 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 sublethal response in polluted waters (Svecevičius, 

2001). However, no studies have been located specifically examining avoidance behavior for Pb 

in fish. Environmental gradients, such as light and pH, can alter preference-avoidance responses. 
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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 aquatic biota to Pb. Physiological effects of Pb on 

aquatic biota can occur at the biochemical, cellular, 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 affinities, and absorption rates. 

Biochemical Effects

Lead was observed to have a gender-selective effect on brain endocannabinoid (eCB) 

(e.g., 2-arachidonylglycerol [2-AG] and N-arachidonylethanolamine [AEA]) levels in fathead 

minnow Pimephales promelas (Rademacher et al., 2005). Cannabinoids, 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 (Sañudo-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 significantly higher levels of AEA and 2-AG compared to males. 

At a concentration of 1000 μg Pb/L, this pattern reversed, with males showing significantly 

higher levels of AEA in the brain than females (Rademacher et al., 2005). After 14-days 

exposure to the 1000 μg Pb/L treatment, 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 (EC50) (Kutlu and 

Susuz, 2004). The exact concentration used in the study was not reported. The EROD enzyme 

is required to catalyze the conjugation 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 bioindicator/biomarker of Pb stress (Blasco and 

Puppo, 1999). A negative correlation was observed 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 amphibians 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 specifically 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 exposure to Pb has also been examined in carp 

(Nakagawa et al., 1995a). After 2-week exposure 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 weeks and placed in a Pb-free environment 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 erythrocyte, 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 counts 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 lymphocytes, 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 protein, 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 exposed 
 AX7-161 

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 Φg 

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 exposure 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 bioindicators of Pb stress in carp (Cyprinus carpio) exposed to 

Pb-nitrates at concentrations of 800 and 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., Ca2+, Mg2+, Na+

, H+, Pb2+) 

(MacDonald et al., 2002; Rogers and Wood, 2003, 2004). Studies suggest that Pb2+ is an 

antagonist of Ca2+ uptake (Rogers and Wood, 2003, 2004). MacDonald et al. (2002) proposed a 

gill-Pb binding model that assumes Pb2+ 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 observed 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 coast 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 concentrations of 0.09 and 0.01 μg/g were associated with 

abalone shell lengths of 3.9 cm (0 years old) and 15.3 cm (8 years old), respectively (Arai et al., 

2002). The authors theorized that young abalones, experiencing rapid growth, do not 

discriminate between the uptake of essential and nonessential metals. However, as abalones 

grow larger and their rate of growth decreases, they increasingly favor the uptake of essential 
 AX7-162 

metals over nonessential metals. This is demonstrated by the relatively consistent concentrations 

of Cu, Mn, and Zn that were reported for 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, interfered with respiration and resulted in death by 

anoxia (Aronson, 1971; National Research Council of Canada., 1973). 

AX7.2.3.4 Factors That Modify Organism Response to Lead 

A great deal of research has been undertaken recently to better understand the factors that 

modify aquatic organism response to metals including lead. A discussion of research on the 

many factors that can modify aquatic organism response 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, controlled by the physiological 

conditions of an organism, including such factors as size, age, point in reproductive cycle, 

nutrition, and overall health. Of these physiological 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 bioaccumulation 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 PbCl2 for 4 days. Canli and Furness 

(1993) found similar results in the Norway lobster Nephrops norvegicus exposed to 100 μg/L of 

Pb(NO3)2 for 30 days. No significant 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 and 
 AX7-163 

sometimes dominate the total Pb accumulated (Knowlton 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 decrease rapidly with increased 

body size and then stabilize) (MacLean et al., 1996). 

Drava et al. (2004) investigated Pb concentrations in the muscle of red shrimp Aristeus 

antennatus from the northwest Mediterranean. Lead 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 relationship between age, shell length and Pb 

concentrations in muscle tissue. The relationship was consistent despite habitat variations in 

Pb concentrations between the study sites, suggesting that Pb concentrations 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 during two consecutive springs 

from three Derbyshire rivers. Results indicated that Pb burden increased slightly with age. 

Similarly, Köck et al. (1996) found that concentrations of Pb in the liver and kidneys of Arctic 

char (Salvelinus alpinus) taken from oligotrophic alpine lakes were positively 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 contrast, Farkas et al. (2003) 

found a negative relationship between Pb concentrations 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 waters 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 route of uptake and feeding 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 environment-related variables (e.g., 

exposure) (Farkas et al., 2003). 

Genetics 

There are few studies documenting the effects 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 effects 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 populations to better cope with the 

chemical; or (2) a contaminant can be genotoxic, meaning it can produce alterations in nucleic 

acids at sublethal exposure concentrations, resulting in changes in hereditary characteristics or 

DNA inactivation (Shugart, 1995). Laboratory studies have shown that exposure to Pb2+ at 

10 mg/mL in blood produces chromosomal aberrations (i.e., deviations in the normal structure or 

number of chromosomes) in some organisms (Cestari et al., 2004). Effects 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 natural 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 individuals have a selective advantage over vulnerable 

individuals in polluted environments, 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 invertebrates. These changes have resulted in decreased 

genetic diversity and are thought to be a potential source of population instability (Duan et al., 

2000; Kim et al., 2003). 

Kim et al. (2003) investigated genetic differences and population structuring in the 

gastropod Littorina brevicula from heavy-metal polluted and unpolluted environments. 

Organisms from polluted sites contained a mean of 1.76 μg Pb/g, while organisms from 
 AX7-165 

unpolluted sites contained 0.33 μg Pb/g. They found significant differences in haplotypes 

between the test groups and allelic diversity was significantly lower among L. brevicula from 

polluted regions. In contrast, Yap et al. (2004) performed a similar 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 metals, exhibited a higher percentage of polymorphic 

loci and excess heterozygosity compared to those from uncontaminated sites. The higher level 

of genetic diversity was attributed to greater environmental heterogeneity (i.e., variation due to 

pollution gradients) in contaminated sites (Yap et al., 2004). 

Duan et al. (2000) investigated amphipod (Hyalella azteca) selective mortality and 

genetic structure following acute exposure to Pb (5.47 mg/L Pb(NO2)2) as well as exposure to 

other heavy metals. They found that genetic differentiation consistently increased among 

survivors from the original population, supporting the hypothesis that heavy metals, including 

Pb, have the potential to alter the 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 higher concentrations (up to 

5000 μg/L) did not result in significant breakage 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 reached. 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 kidney cell cultures, although when assessed at 

the end of the longer exposure periods, aberrations were less common. 

Environmental Biological Factors 

Environmental factors that are biological in origin can alter the availability, uptake and 

toxicity of Pb to aquatic organisms. These factors can be grouped into living and non-living 

constituents. For example, living organisms may sequester Pb from the water column, reducing 

the availability and toxicity of the metal in the water column to other biota, thus reducing 
 AX7-166 

potential toxic effects in other organisms. Non-living organic material (e.g., components of 

sloughed-off scales, mucus, carcasses, and other decomposing, humic material) can similarly 

combine with Pb from the water column, rendering it unavailable. This section will review the 

literature on biological environmental factors and their influence on the bioavailability, uptake, 

and toxicity of Pb. 

Van Hattum et al. (1996) studied the influence of abiotic variables, including DOC on Pb 

concentrations in freshwater isopods (Proasellus meridianus and Asellus aquaticus). They found 

that BCFs were significantly negatively correlated 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) investigated the combined effects of Pb and humic acid on total 

chlorophyll content, growth rate, multiplication rate, and Pb uptake of common duckweed. 

When humic acid was added to the Pb-nitrate test solutions (50, 100, and 200 mg Pb(NO3)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, resulting 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 effects as they related to the optical 

properties of the various NOM samples. The results showed that NOM in test water almost 

always increased LT50 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 biological origin in the environment have been 

shown to reduce Pb availability and, therefore, toxicity in some aquatic organisms. It is 

generally thought that this occurs through complexation or chelation processes that take place in 

the water column. 
 AX7-167 

Physicochemical Environmental Factors 

This section reviews the literature on physicochemical environmental factors and their 

influence on the bioavailability, uptake, and toxicity of Pb in aquatic organisms. These factors 

are discussed with regard to their influence 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 aquatic organisms (Weber, 1993). Significantly lower survival, decreased 

hatching success, slower development, and increased egg mass and larval 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 period, mortality increased 

with decreasing pH for the bivalve Pisidium casertanum, while pH-independent mortality was 

reported for gastropods and crustacea under similar 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-Berigan 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., 1993). Lead was also more toxic to fathead 

minnows at lower pH levels (Schubauer-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 sediments (Lewis and McIntosh, 1986). In A. aquaticus, 

temperature increases were found to be more important 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., 2004). Lead and zinc 

body concentrations in Asellus sp. were found to vary markedly with seasonal temperature 

changes, with greater concentrations present in spring and summer (Van Hattum et al., 1996). 

Acute and chronic toxicity of Pb increases with decreasing water hardness, as Pb becomes 

more soluble and bioavailable to aquatic organisms (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 exposed to Pb for 28 days at low pH and low water 

hardness experienced 50% mortality, while exposure to Pb at high pH and high water hardness 
 AX7-168 

resulted in 91.7% survival (Horne and Dunson, 1995a). Exposure to Pb at high pH and low 

water hardness or low pH and high water hardness resulted in 75 and 41.7% survival, 

respectively (Horne and Dunson, 1995a). Similar results were reported for Jefferson 

salamanders during a 7-day exposure and wood frogs during 7- and 28-day exposures (Horne 

and Dunson, 1995c). In some cases, water hardness 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 Ca2+ concentrations have been shown to protect 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 Ca2+ concentrations decrease, the passive flux of ions (e.g., Pb) and 

water increases. At the lowest waterborne Ca2+ 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 Ca2+ concentrations, Pb 

accumulation did not significantly increase with Pb concentration in water. This result 

demonstrates the protective effects of waterborne Ca2+ and supports the suggestion that the Ca2+

component of water hardness determines the toxicity of Pb to fish (Rogers and Wood, 2004). 

Rogers and Wood (2004) reported that the uptake of Ca2+ and Pb2+ involves competitive 

inhibition of apical entry at lanthanum-sensitive Ca2+ channels and interference with the function 

of the ATP-driven baso-lateral Ca2+ pump. High mortality was reported in brown trout (Salmo 

trutta) fry exposed to Pb at a waterborne Ca2+ concentration of 20 μmol/L, while negligible 

mortality was reported at the same Pb concentration but at a waterborne Ca2+ concentration of 

200 μmol/L (Sayer et al., 1989). Adverse effects to mineral uptake and 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 sulfide (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 a 
 AX7-169 

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 modifying factor to metal toxicity. 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 LC50 for Pb was 1140 μg/L 

(95% CL = 840, 1440 μg/L). At an increased salinity of 25l, the toxicity of Pb was substantially 

reduced (LC50 = 4274 μg/L [95% CL = 3540, 5710 μg/L]) (Verslycke et al., 2003). The 

reduction in toxicity was attributed to increased complexation of Pb2+ with Cl!

 ions. 

Nutritional Factors 

The relationship between nutrition and Pb toxicity 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 nutrition (the process by which 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 metabolic 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 reducing growth inhibition in blue-green algae 

(Synechococcus aeruginosus) exposed to 200 mg Pb/L. Exposure 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 from 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 significant impact on Pb toxicity in S. aeruginosus. In a 

similar study by Rao and Reddy (1985) on Scenedesmus incrassatulus, nitrogen, phosphate and 
 AX7-170 

carbon sources (including glucose), all had protective effects, and reduced Pb toxicity at 300 and 

400 mg Pb/L. Both studies proposed similar hypotheses regarding nutrient-Pb mechanisms that 

led to reduced toxicity. One hypothesis was that the nutrients were able to reverse toxic effects. 

The second hypothesis was that the nutrients interacted 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) investigated 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. Ammonium uptake was inhibited, and similarly, 

inhibition decreased from 72% inhibition after 24 h to 26% inhibition after 96 h of exposure. 

Carbon dioxide fixation and nitrogenase activity followed similar patterns, and results indicated 

that Pb exposure can affect the uptake of some nutrients in N. muscorum. 

Adam and Abdel-Basset (1990) studied the effect of Pb on metabolic processes of 

Scenedesmus obliquus. They found that nitrogenase activity was inhibited by Pb nitrate, but 

enhanced by Pb-acetate. As photosynthetic products 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 increase 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 lower Pb concentrations (<0.207 mg/L) and increased nutrient 

concentrations, algae release more exudates that form inert complexes 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 consistent with those of 

Capelo et al. (1993), who investigated uptake of nitrogen and phosphorus in the algae 

Selenastrum capricornutum over time in the absence and presence of 0.207 mg Pb/L. They 

found that the presence of Pb had 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 hypothesized that this relationship was due in part to a 

diluting effect of the food. 

In summary, nutrients affect Pb toxicity in those 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 nutrient uptake ability, though it 

has been demonstrated that Pb exposure may lead to increased production and loss of organic 

material (e.g., mucus and other complex organic ligands) (Capelo 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 uptake and exert toxicity. Research on the interactions of Pb with 

complexing ligands and other physical and biological factors was more thoroughly discussed in 

Section AX7.2.3.4. Predicting the response of organisms 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 antagonistic). The mode of action model assumes 

that each of the individual substances has similar pharmacokinetics as well as mode of action. 

Approaches to modeling combined mode of action 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 additivity. 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 chemical 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 suggests that the level of 

effect of a mixture will be the sum of the effects of each chemical used separately. Thus, the 
 AX7-172 

separate concentration-effects models for each chemical 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 additive (antagonistic) or more than 

additive (synergistic) (see Piegorsch et al., 1988; Finney 1947 for further information). 

The complexity of metal mixture interactions 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 toxicity 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 Pb. For example, Pb is a wellknown antagonist to Ca2+ (Niyogi and Wood, 2004; Hassler et al., 2004), which is an essential 

element, required for a number of physiological processes in most organisms. Lead ions have an 

atomic structure similar to Ca2+ and can be transported either actively or passively across cell 

membranes in place of Ca2+. An example of this interaction was reported by Behra (1993a,b) 

where Pb was shown to activate calmodulin reactions in rainbow trout (O. mykiss) and sea 

mussel (Mytilus sp.) tissues in the absence of calcium. Calmodulin (CaM) is a major 

intracellular calcium receptor and regulates the 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 required 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 Mg2+, Na+

, and Cl!

regulation in fish (Ahern and Morris, 1998; Rogers and Wood, 2003, 2004; Niyogi and Wood, 

2004). Li et al. (2004) reported on the interaction of Pb2+ with Cd2+ in the context of adsorption 

from solution by Phanerochaete chrysosporium, a filamentous fungus. The authors found that 

cadmium uptake decreased with increasing concentration of Pb ions with Pb2+ outcompeting 

Cd2+ for binding sites. 

Hassler et al. (2004) reported that in the presence of copper (Cu2+), there was a 

significantly higher rate of internalization of Pb in the green algae Chlorella kesserii. It was 
 AX7-173 

suggested that Cu2+ may have affected organism physiology through the disruption of cell 

membrane integrity. This would allow increased cation (i.e., Pb2+) permeability and, therefore, 

substantially increase internalization of Pb and result in effects that were more than additive 

(synergistic). Hagopian-Schlekat et al. (2001) examined 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-spiked sediment and pore water. The mixed 

metal sediment toxicity tests demonstrated 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 concentrations, (2:1 Pb:Zn ratio), 

a synergistic toxic effect was observed in the frog larvae relative to the effects observed from 

exposure to the individual metals and at higher zinc concentrations. Enhanced Pb toxicity was 

attributed to the interference of Pb with cellular activities due to binding with sulfhydryl 

polypeptides and nucleic acid phosphates (Herkovits and Perez-Coll, 1991). Allen (1994) 

reported on the accumulation of numerous metals and ions into specific tissues of the tilapia 

Oreochromis aureus. Tilapia exposed to low concentrations 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 concentrations of cadmium with low concentrations of Pb 

caused increased uptake of Pb in certain organs (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 metals, including Pb, 

and a combined metal mixture under changing salinity conditions. At a salinity of 5%, the 

reported LC50 for Pb was 1140 μg/L (840, 1440 μg/L). At an increased salinity of 25l, the 

toxicity of Pb was substantially reduced (LC50 = 4274 μg/L [3540, 5710 Φg/L]) (Verslycke 

et al., 2003). This reduction in toxicity was attributed to the increased concentration of Cl!

 ion 

due to increased salinity, in that it complexed with divalent Pb in the test 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 toxic effects of contaminant mixtures (metals and 

organics) in sediments. This approach has been used in numerous surveys and studies since 
 AX7-174 

1994. mSQGQs are useful to risk assessors but their inherent limitations and underlying 

assumptions must be fully understood (see Section AX7.2.1.5). 

Summary of Interactions With Other Pollutants 

Norwood et al. (2003) reported that in a review and reinterpretation 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 interaction 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 influence on generation of ground-level ozone, 

depletion of stratospheric ozone, global warming, or other indicators 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 ground-level nitrogen oxides or volatile organic 

compounds, thus precluding generation of ground-level ozone. 

Approached from another viewpoint, climate change can have a major impact on the 

fate/behavior of Pb in the environment and, therefore, 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 of 
 AX7-175 

factors that influence Pb accumulation in aquatic organisms, mechanisms of detoxification and 

avoidance of Pb, and greater understanding of the interactions of Pb in aquatic systems. All of 

these factors require some consideration when attempting 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 critical recent advancements in the understanding 

of the toxicity of Pb to aquatic biota. The section begins with a review of the major findings and 

conclusions from the 1986 Lead AQCD (U.S. Environmental Protection Agency, 1986a). The 

following sections summarize the research conducted since 1986 on determining the 

concentrations of Pb that cause the 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 materials in controlled studies to permit evaluation 

of specific impacts on higher levels of organization (beyond the individual organism). Potential 

effects at the population level or higher are, of necessity, extrapolated from individual level 

studies. Available population, community, or ecosystem 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). Therefore, the best documented links 

between lead and effects on the environment are with 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 Protection Agency, 1986a) reviewed data in 

the context of the sublethal effects of lead exposure. The document 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 comprehensive 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 algae, 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 concentrations 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 inhibition of the enzyme 

ALAD, required for hemoglobin synthesis. At high 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 aquatic invertebrates in the 1986 

Lead AQCD (U.S. Environmental Protection Agency, 1986a). In general, lead concentrations in 

aquatic invertebrates were found to be correlated closely with concentrations in water rather than 

food. Freshwater snails were found to accumulate lead in soft tissue, often 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. Environmental 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. Environmental Protection Agency, 1985). 
 AX7-177 

AX7.2.4.2 Recent Studies on Effects of Lead on Primary Producers 

Using literature published since 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 processes (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, 1990; 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. Effects 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 calculating the LOEC, NOEC, or EC50 values. Clinical signs of Pb toxicity include 

the deformation and disintegration of algae cells 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 cell 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 growth that was 52.6 and 17.4%, respectively, 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 (Pb2+) 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 EC50 for growth inhibition at 

13,180 μg/L (95% CI: 10,190, 14,620). Decreased cell number, but increased cell size, was 
 AX7-178 

observed in Selenastrum capricornutum7 exposed to lead (Pb2+) 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 healthy (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, and 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 inhibition 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-acetate 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-nitrate or Pb-acetate resulted in cell numbers 

that were 81 and 90% of the controls, respectively (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 alone (Rai and Raizada, 1989). Nostoc muscorum exposed to 

chromium and Pb in combination demonstrated better growth than when exposed to either of the 

metals alone (Rai and Raizada, 1989). Antagonistic interaction was observed in the exposure of 

Nostoc muscorum to Pb and nickel in combination (Rai and Raizada, 1989). When applied 

separately, these metals demonstrated different levels of toxicity; 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 plant growth has been studied using Spirodela polyrhiza, 

Azolla pinnata, and Lemna gibba (Gaur et al., 1994; Gupta and Chandra, 1994; Miranda and 

Ilangovan, 1996). Test durations ranged from 4 to 25 days and test concentrations 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 growth and chlorophyll and 

protein content. 

Of the species reviewed here, the effects of Pb on aquatic plant growth are most 

pronounced in Azolla pinnata (Gaur et al., 1994). An EC50 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 EC50 for growth of 3730 μg/L (Gaur et al., 1994). Lemna gibba

was shown to be the least sensitive plant species to Pb: significant growth inhibition was 

reported at concentrations of 200,000 μg/L or greater after 25 days of exposure to concentrations 

of 30,000, 50,000, 100,000, 200,000, 300,000, or 500,000 μg/L (Miranda 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 test concentrations (Miranda and Ilangovan, 1996). 

Clinical signs of Pb toxicity include yellowing and disintegration of fronds, reduced frond size, 

and chlorosis (Gaur et al., 1994; Miranda and Ilangovan, 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 are variously affected by exposure to Pb, both 

singularly and in combination with other metals. Lead adversely affects the metabolic processes 

of nitrate uptake, nitrogen fixation, ammonium uptake, and carbon fixation at concentrations of 

20,000 μg Pb/L or greater (Rai and Raizada, 1989). Lead in combination with nickel has an 

antagonistic effect on nitrogen fixation and ammonium uptake, but a synergistic effect on nitrate 

uptake and carbon fixation (Rai and Raizada, 1989). Lead in combination with chromium has an 

antagonistic effect on nitrate uptake, but it has a synergistic effect on nitrogen fixation, 

ammonium uptake, and carbon fixation (Rai and Raizada, 1989). 

Lead effects on nitrate uptake in Nostoc muscorum (μg NO3/μg Chl a) were greatest after 

24 h, when exposure to 20,000 μg/L reduced nitrate uptake by 64.3% compared to controls. 
 AX7-180 

Nitrate uptake reported after 48, 72, and 96 h was reduced by 30.0, 37.5, and 38.9%, 

respectively, compared to controls (Rai and Raizada, 1989). Lead in combination with 

chromium, both at a test concentration of 20,000 μg/L, demonstrated antagonistic effects on 

nitrate uptake. Compared to controls, nitrate uptake was reduced by 52.4, 30, 25, and 22.2% at 

24, 48, 72 and 96 h, respectively (Rai and Raizada, 1989). The greatest effect on uptake 

occurred at 24 h when, compared to controls, a 52.4% reduction was reported in the test 

concentration. Lead and nickel in combination at test concentrations of 20,000 and 1000 μg/L, 

respectively, resulted in a greater 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 C2H4/ 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 concentration 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 controls, nitrogenase activity was reduced by 42.9, 32.7, 

and 13.6% at 24, 48, and 72 h, respectively (Rai and Raizada, 1989). Lead and chromium, both 

administered at a concentration of 20,000 μg/L, had a synergistic impact on nitrogenase activity 

in Nostoc muscorum. Nitrogenase activity in the test 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 NH4 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 ammonium 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, respectively, 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, respectively (Rai and Raizada, 1989). Lead in 

combination with chromium, both at concentrations 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 experienced the greatest reduction in carbon 

fixation at 0.5 h of exposure: 62% compared to controls. Inhibition of carbon fixation in the test 

group was less pronounced after 1 and 2 h of exposure: 29 and 13% of controls (Rai and 

Raizada, 1989). Lead in combination with nickel 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, and 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 effects 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 (K2HPO4 and Na2HPO4) 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 was 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), EC50 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 concentrations >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 primary producers experience EC50 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 effects 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 adversely affects the metabolic processes of nitrate 

uptake, nitrogen fixation, ammonium uptake, and carbon fixation (Rai and Raizada, 1989). Lead 

in combination with nickel or chromium produced synergistic effects for nitrate uptake, 

nitrogenase activities, ammonium uptake, and carbon fixation (Rai and Raizada, 1989). 

Leads Effects on Primary Productivity 

Lead nitrate and Pb-acetate have been shown to have adverse effects on the primary 

productivity of aquatic plants in two water bodies in India (Jayaraj et al., 1992). One of the two 

water bodies was a freshwater tank that receives wastewater and supports a rich population of 

hyacinths, and the other was a wastewater stabilization pond. Water quality characteristics in the 

freshwater tank were pH = 7.5, dissolved oxygen = 6 mg/L, and water hardness (CaCO3) = 

100 mg/L. Water quality characteristics in the wastewater pond were pH = 8.1, dissolved 

oxygen = 6.2 mg/L, and water hardness (CaCO3) = 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 suspended in each of the water bodies for 4 h. The 

concentrations of Pb-acetate (5000, 10,000, 25,000, 50,000, and 100,000 μg/L) were applied in 

the same manner. The EC50 values were determined based on the concentration 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 EC50 values for GP and NP were 25,100 and 6310 μg/L, 

respectively, compared to Pb-acetate EC50 values of 50,100 and 28,200 μg/L for GP and NP, 

respectively (Jayaraj et al., 1992). In the stabilization pond, Pb-nitrate EC50 values for GP and 

NP were 31,600 and 28,200 μg/L, respectively, compared to Pb-acetate EC50 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 attributed to differences in species composition and diversity. The 

freshwater tank was dominated by water hyacinths that decreased the photic zone available for 
 AX7-183 

photosynthesis and consumed a great deal of available nutrients. The stabilization pond had 

a rich nutrient budget, resulting in improved alga growth and species diversity (Jayaraj 

et al., 1992). 

AX7.2.4.3 Recent Studies on Effects of Lead on Consumers 

This section focuses on the effects of Pb to aquatic biota including invertebrates, 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 concentrations 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 numerous electronic bibliographic 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 literature search and recent reviews of the toxicity of Pb (Eisler, 

2000), numerous studies have been published on the toxicity of Pb to aquatic consumers. 

Hardness, pH, temperature, and other factors are important considerations when characterizing 

the acute and chronic toxicity 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 comparing effects between studies difficult. Studies 

reporting only physiological responses to Pb exposure (e.g., reduction 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 invertebrates, 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 following presents information on the toxicity of 

Pb to invertebrates in fresh and marine waters. 
 AX7-184 

Freshwater Invertebrates 

Acute and chronic Pb toxicity data for freshwater invertebrates are summarized in Table 

AX7-2.4.1. As described in Section AX7.2.3.5, water hardness is a critical factor governing the 

solubility, bioavailability, and ultimately 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 

CaCO3/L; hard water, 124 mg CaCO3/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 (CaCO3) = 71 mg/L) mortality was >50% 

for H. azteca at a dissolved Pb concentration of 151 μg/L. The LOEC for survival in hard water 

(hardness (CaCO3) = 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 the 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 increased 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 certain life stages may be important when 

determining the toxicity of metals to test organisms. For example, Bodar et al. (1989) exposed 

early life stages of Daphnia magna to concentrations of Pb(NO3)2. The test medium had a pH 

of 8.3 ∀ 0.2, water hardness (CaCO3) of 150 mg/L, and temperature of 20 ∀ 1 ΕC. Lead 
 AX7-185 Table AX7-2.4.1. Effects of Lead to Freshwater and Marine Invertebrates Species Chemical Endpoint: Conc. (μg/L)* Duration of Exposure Water Chemistry Test Type - Effe

Freshwater Cladoceran (Ceriodaphnia dubia) lead chloride LC50: 280 >2,700 >2,700 48 h pH: 6–6.5 7–7.5 8–8.5 Hardness: 280-300 mg/L CaCO3 static-survival 

Worm (Lumbriculus variegatus) 

lead chloride 

LC50: 

>8,000 >8,000 >8,000 

96 h pH: 

6–6.5 7–7.5 8–8.5 Hardness: 280-300 mg/L CaCO3

static-survival 

Amphipod (Hyalella azteca)

lead chloride 

LC50: 

<90 >5,400 >5,400 

96 h pH: 

6–6.5 7–7.5 8–8.5 Hardness: 280-300 mg/L CaCO3

static-survival 

Amphipod (Hyalella azteca) 

lead chloride 

LC50: 

27 (20.1-36.4) 

8 days Hardness 

130 mg/L pH 7.8-8.6 

renewal, 1-weekold amphipods 

Amphipod (Hyalella azteca) 

lead chloride 

LC50: 

60 (53.6-67.3) 

8 days Hardness 

130 mg/L pH 7.8–8.6 

renewal, 10- to 1week old amphipods 

Mayfly (Leptophlebia marginata)

lead chloride 

LC50: 

1090 (400-133200) 

96 h pH: 

4.5 

acute - survival 

Mayfly (Leptophlebia marginata) 

lead chloride 

LC50: 

5000 

96 h pH 

7.0 

acute - survival 
 AX7-186 Table AX7-2.4.1 (cont’d). Effects of Lead to Freshwater and Marine Invertebrates Species Chemical Endpoint: Conc. (μg/L)* Duration of Exposure Water Chemistry Test Type - Effe

Amphipod (Hyalella azteca) 

lead nitrate 

LC50: 

10 21 18 

96 h pH: 

5.0 5.5 6.0 

acute-survival 

Bivalve (Pisidium compressum) 

lead nitrate 

LC50: 

38,000 21,300 11,400 

96 h pH: 

3.5 4.0 4.5 

acute- survival 

Bivalve (Pisidium casertanum) 

lead nitrate 

LC50: 

23,600 23,500 56,000 

96 h pH: 

3.5 4.0 4.5 

acute- survival 

Gastropod (Amnicola limosa)

lead nitrate 

LC50: 

10,300 20,600 9,500 

96 h pH: 

3.5 4.0 4.5 

acute- survival 

Mussel (Dreissena polymorpha) 

lead nitrate 

EC50: 

370 91 

48 h 

10 wks 

pH = 7.9; Hardness = 150 mg CaCO3/L; Temp = 15 ΕC 

renewal - filtratio

Mussel (Dreissena polymorpha) 

lead nitrate 

LT50: 

358 

72 days pH = 7.9; Hardness = 150 mg 

CaCO3/L; Temp = 15 ΕC 

renewal - filtratio

Amphipod (Hyalella azteca) 

lead nitrate 

LC50: 

4.8 (3.3 - 7.1) 

7 days pH = 7.37 - 8.27 

Hardness = 18 mg CaCO3/L 

DOC = 0.28 mg/L 

renewal - surviva

Amphipod (Hyalella azteca) 

lead nitrate 

LC50: 

113 (101 -126) 

7 days pH = 8.21 - 8.46 

Hardness = 124 mg CaCO3/L; 

DOC = 1.1 mg/L 

renewal - surviva
 AX7-187 Table AX7-2.4.1 (cont’d). Effects of Lead to Freshwater and Marine Invertebrates Species Chemical Endpoint: Conc. (

Φg/L)* 

Duration of 

Exposure Water Chemistry Test Type - Effe

Mayfly (Leptophlebia marginata) 

lead chloride 

LC50: 

1090 (95% C.I. = 133.2) >5000 

96 h pH = 4.5 - 6.5; 

DOC - 21.6 mg Cl!1; 

Cond = 7.0 

ΦS cm!1

renewal - surviva

Cladoceran (D. magna) 

lead nitrate 

LC50: 

0.45 

48 h pH = 8.3

∀ 0.2 

Hardness (CaCO3) = 150 mg/L 

Temp= 20 ΕC 

static - embryogenesis 

Cladoceran (D. magna) 

lead chloride 

NOEC: 260 

12 to 21 d Not specified renewal - 

reproduction 

Cladoceran (D.magna) 

lead chloride 

NOEC: 270 

10 d Not specified renewal - growth 

Amphipod (Hyalella azteca) 

lead LOEC: 

(Dissolved Pb) 192 (Total Pb) 466 

96 h pH = 8.27 

Hardness (CaCO3) = 275 mg/L 

Temp = 21.1 ΕC 

flow through - survival 

Tubificid worm (Tubifex tubifex) 

lead nitrate 

EC50: 

237 (183–316) 142 (107–184) 

24 h 48 h 

pH = 7.5–7.7 Hardness = 245 mg/L Temp = 29.5–31 ΕC 

static - immobilization 

MarineCopepod (Amphiascus tenuiremis) 

lead LC50: sediment 

2462 μg metal/dry sediment 

96 h pH = 7.7 

∀ 0.1 

Dissolved O2 -6.3 

∀ 0.3 mg/L 

Salinity – 32 ppt 

 

Bivalve (Mytilus galloprovincialis) 

lead nitrate 

EC50 : 221 

(58.9–346.3) LOEC : 50 

 artificial seawater embryogenesis 

* - 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 chlorion layer. The chlorion layer in other species (e.g., rainbow 

trout) is known to adsorb metals, thereby, preventing 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 exposure 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 important modifying factor to Pb toxicity. 

Verslycke et al. (2003) exposed the estuarine mysid Neomysis integer to individual metals, 

including Pb, and metal mixtures under changing salinity. Water temperature (20 ∀ 1o

C) and 

salinity were reported, although no other water quality parameters were available (e.g., pH, water 

hardness). At a salinity of 5l, the reported LC50 for Pb was 1140 μg/L (95% CI: 840, 

1440 μg/L). At an increased salinity of 25l, the toxicity of lead was substantially reduced (LC50

= 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 toxicity of Pb-chloride in sediment and sediment 

pore water to female and male estuarine copepods Amphiascus tenuiremis. The reported LC50

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 sensitive (p = 0.038) to Pb than females as 

determined by generalized linear model analysis. 

Beiras and Albentosa (2003) examined the inhibition of embryo development in 

commercial bivalves Ruditapes decussatus and Mytilus galloprovincialis after exposure to 

concentrations of Pb(NO3)2 in seawater. No water chemistry parameters other than temperature 

were reported (test conducted at 20 oC). An EC50 range for R. decussatus was reported as 156 to 

312 μg/L, as insufficient data were available to calculate the actual EC50. The lowest observable 

effect concentration (LOEC) was 156 μg/L. For M. galloprovincialis, the EC50 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 toxicity in fish include production of excess mucus, lordosis, 

anemia, darkening of the dorsal tail region, degeneration of the caudal fin, destruction of spinal 

neurons, ALAD inhibition, growth inhibition, renal pathology, reproductive effects, growth 

inhibition, and mortality (Eisler, 2000). Toxicity in fish has 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 waters. Table AX7-2.4.2 summarizes 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 subsequent toxicity have been well-studied (Sayer 

et al., 1989; Spry and Wiener, 1991; Schubauer-Berigan et al., 1993; Stouthart et al., 1994; 

MacDonald et al., 2002; Rogers and Wood, 2003). Schubauer-Berigan et al. (1993) exposed 

fathead minnow to Pb-chloride over 96 hours. The reported LC50 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 water 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 concentration (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 reproductive abnormalities were noted in mature 

male rainbow trout at 10 μg/L Pb-nitrate (Ruby et al., 1993). 

Schwartz et al. (2004) examined 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 ΕC, and Pb was added as PbCl2. NOM from a 
 AX7-190 Table AX7-2.4.2. Effects of Pb to Freshwater and Marine Fish Species Chemical Endpoint: Conc. (

Φg/L) 

Duration of 

Exposure Water Chemistry Comments 

Freshwater 

Fathead minnow (Pimephales promelas) 

lead chloride 

LC50: 

810 >5,400 >5,400 

96 h pH: 

6–6.5 7–7.5 8–8.5 Hardness: 280-300 mg/L CaCO3

static, measured 

Rainbow trout - mature males (Oncorhynchus mykiss) 

lead nitrate 

Reproductive effects: 10 

12 days Hardness 128 mg/L 

CaCO3

Decreased spermatocyte development 

Fathead minnow (Pimephales promelas) 

lead acetate 

Reproductive Effects: 500 

29 days pH: 

7.5–8.5; Hardness 130 mg/L CaCO3; 

22–25 ΕC 

(Pb 95% soluble) 

Fewer viable eggs produced, testicular damage 

Rainbow trout – Juvenile (Oncorhynchus mykiss) 

lead nitrate 

LC50: 1000 

(800 - 1400) 

96 h pH: 

7.9–8.0 DOC = 3 mg/L Hardness (CaCO3) = 

140 mg/L 

Flow through - Surviv

Common carp (Cyprinus carpio) 

not reported 

LC50: 

6.5 cm fish – 1030 3.5 cm fish – 300 

96 h pH: 

7.1 Temperature–15 ΕC 

Oxygen sat. 6.4 mg/L 

static-renewal - Survival 
 AX7-191

number of U.S. rivers and lakes was then added to the test system, and the LT50 was reported. 

NOM was found to reduce the toxic effects of Pb to rainbow trout. 

Fish size is an important variable in determining 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 mortality. Water chemistry parameters were reported 

(pH = 7.1; temperature = 20 ΕC). Smaller fish (3.5 cm) were found to be more sensitive to Pb 

than were larger fish (6.5 cm). The reported LC50s were 0.44 mg/L and 1.03 mg/L, respectively. 

Marine Fish 

There were no studies available that examined the toxicity of Pb to marine fish species for 

the time period examined (1986 to present). However, Eisler (2000) reviewed 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). Organolead compounds (e.g., tetramethyl Pb, tetraethyl Pb, 

triethyl Pb, diethyl Pb) were generally more toxic to plaice than inorganic Pb (Maddock and 

Taylor, 1980). 

Other Aquatic Biota 

A paucity of data exist on the effects of Pb to growth, reproduction, 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 water hardness of 233 to 244 mg CaCO3/L, pH from 7.85 to 7.9, 

and temperature at 23 ΕC. Frog larvae were found to display little to no activity in the low 

oxygen and high Pb treatment. Hypoxia-like behavior 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 and 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 longer periods (>4 weeks) were smaller and 

metamorphosed later compared to unexposed individuals. 
 AX7-192 Table AX7-2.4.3. Nonlethal Effects in Amphibians Species Chemical Endpoint: Concentration Duration of Exposure Water Chemistry Comments 

Frogs (Rana ridibunda) 

lead nitrate Biochemical effects: 

14,000 μg/L 

30 days Not specified Hepatic ALAD decreased by 

90% 

Frogs (Bufo arenarum) 

 Mortality: 

16 mg Pb2+/L 

5 days Not specified Effects reported include erratic 

swimming, loss of equilibrium 

Frogs (Rana catesbeiana) 

 Hypoxia-like behavior:

780 μg/L 

7 days O2 = 3.5-7.85 mg/L 

pH = 7.85-7.9 Temp = 23

ΕC 

CaCO3 = 233-244 

mg/L 

Larvae used 

Turtle Hatchlings (Trachemys scripta) 

lead acetate NOEL: 100 μg/g 

(Survival and behavior) 

4 weeks N/A Exposure via single injection 
 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 Pb2+/L 

and 16 mg Pb2+/L. Relative to controls, the 8 mg Pb2+/L treatment group exhibited 40% 

mortality and the 16 mg Pb2+/L group 60% mortality after 120 h (p < 0.05). The authors reported 

behavioral effects, erratic swimming, 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 turnover, seeking cover), growth, and survival of hatchling 

slider turtles (Trachemys scripta) exposed to Pb-acetate were investigated in one study (Burger 

et al., 1998). In the first part of the study, 6-month-old hatchlings 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 study, however, several effects were 

reported from the second part of the study at doses in the range of 250 to 2,500 μg/g bw. As the 

dose increased, so did the plastron length (i.e., ventral section of the shell), carapace length, and 

weight. The highest dose group had the lowest survival rate with an LD50 of 500 μg/g bw. 

Behavioral effects included slower times of righting behavior and seeking cover. The authors 

suggested a NOEL of 100 μg/g bw for slider turtles for survival and behavior. 

AX7.2.4.4 Recent Studies on Effects of Lead on Decomposers 

In this section, decomposers are defined as being bacteria and other microorganisms. 

Many invertebrates are also potentially considered decomposers, but the effects of Pb to 

invertebrates have been described in previous sections. There were 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 observed at b concentrations ranging from 100 to 

200,000 μg/L. Clinical signs of Pb toxicity in plants 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 synthesis, thus affecting photosynthesis, cell cycle and 

division, and ultimately resulting in death. The toxicity of Pb to macrophyte growth has been 
 AX7-194

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 toxicity varying with the species 

and life stage tested, duration of exposure, 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 adverse effects being reported as low as 0.45 μg/L. 

Effects concentrations for aquatic invertebrates 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 quality parameters. Amphibians tend to be relatively Pb tolerant 

(Eisler, 2000); however, they may exhibit decreased 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 effects 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 considered in future reviews of lead toxicity. 

In considering these higher levels of organization, 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 the influences of redundancy/interactions should 

be taken into account. 

AX7.2.5 Effects of Lead on Natural Aquatic Ecosystems 

Introduction 

As discussed previously, 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 materials in controlled studies to 

permit evaluation of specific impacts on higher levels of organization (beyond the individual 

organism). Potential effects at the population level or higher are, of necessity, extrapolated from 

individual level studies. Available population, community, or ecosystem 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). Therefore, 
 AX7-195

the best documented links between lead and effects on the environment are with effects on 

individual organisms. 

This section discusses the effects of Pb on natural aquatic ecosystems. Such effects 

include changes in species composition and richness, ecosystem function, and energy flow due to 

Pb stress. The format of this section generally follows a conceptual framework for discussing 

the effects of a stressor such as Pb on an ecosystem. This conceptual framework was developed 

by the EPA Science Advisory Board (Young and Sanzone, 2002). The essential attributes used 

to describe ecological condition include landscape condition, biotic condition, chemical and 

physical characteristics, ecological processes, hydrology and geomorphology and natural 

disturbance regimes. The majority of the published literature pertaining to Pb and aquatic 

ecosystems focuses on the biotic condition, one of several essential attributes of an ecosystem as 

described in Young and Sanzone (2002). For the biotic condition, the SAB framework identifies 

community extent, community composition, trophic structure, community dynamics, and 

physical structure as factors for assessing ecosystem health. Other factors for assessing the 

biotic condition such as effects of Pb on organs, species, populations, and organism conditions 

(e.g., physiological status) were discussed in Sections AX7.2.3 and AX7.2.4. 

For natural aquatic ecosystems, the focus of study in the general literature has been on 

evaluating ecological stress where the sources of Pb were from urban 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 other studies, although Pb 

deposition was studied, the effects of acid deposition on aquatic life were the 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 phytoplankton 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 observed effect. Studied variables include water 

hardness, pH, temperature, and physical factors such as embeddedness, dominant substrate, and 

velocity. In most cases, single variable statistical techniques were used to evaluate the data. 
 AX7-196

However, in other cases multivariate techniques were used. Therefore, where appropriate, some 

detail on the statistical methods used is presented. 

Although most of the available studies discussed in this section focus on the biotic 

condition, one case study examining multiple components of the EPA conceptual framework is 

also included. The remainder of this 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. Maret et al. (2003) examined several ecological 

components to determine any negative associations with metals and the watershed communities. 

The variables examined and associated ecological conditions are presented in Table AX7-2.5.1. 

In addition to measurements of non-metal variables (e.g., dissolved oxygen levels, water 

temperature and pH, embeddedness), Cd, Pb, and Zn levels were also compared in affected sites 

versus reference sites. 

Some of the above non-metal variables are important to macroinvertebrate communities. 

For example, a stream with highly embedded substrate can have a lower number of individuals 

within a species or a different species composition compared to a stream with less embeddedness 

(Waters, 1995). Macroinvertebrates from the Ephemeroptera (mayflies), Plecoptera (stoneflies), 

and Trichoptera (caddisflies) (EPT) group inhabit the surface of cobble and the interstitial spaces 

between and underneath cobble. When substrate is embedded, these interstitial spaces are filled, 

leaving less habitat space for EPT taxa. In another example, water temperature is important; 

some macroinvertebrates (e.g., stoneflies) are usually only found in cooler water (Harper and 

Stewart, 1984). 

Of the variables examined only metal concentrations, mine density, site elevation, and 

water temperature were significantly 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 explanatory variables. Lead concentrations were 

significantly correlated with the number of mines in proximity to the watershed. Lead 

concentrations in sediment and water were strongly correlated to Pb levels in whole caddisflies, 

r

2

 = 0.90 and 0.63, respectively. Furthermore, mine density was significantly correlated to Pb in 
 AX7-197

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 (km2) 

Production mine density/km2

Biotic Condition Organism condition 

population structure/ 

dynamics 

Caddisfly tissue concentrations (mg/kg) 

Number of EPT taxa 

Density of EPT individuals (no./m2

) 

Chemical/physical 

characteristics 

Chemical/physical 

parameters 

Dissolved oxygen (mg/L) 

Specific conductance (μS/cm) 

Water temperature (E ΕC) 

pH 

Water hardness (mg/L) 

Total NO3 (mg/L) 

Total P (mg/L) 

Dissolved NH3

 (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 (m3

/s) 

Stream width (m) 

Stream depth (m) 

Open canopy (%) 

Stream velocity (m/s) 

Embeddedness (%) 

Dominant substrate (mm) 

Natural disturbance regimes — None measured 

tissue, r2

 = 0.64. Although temperature was significantly different between reference and 

mine-affected sites, temperature conditions were concluded to be non-limiting to aquatic life. 

For example, reference and mine-affected sites had 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 number of metal-sensitive mayfly 
 AX7-198

species was observed. Similar, significant negative correlations were found between sediment 

Pb levels (132 to 6252 μg/g) and the same macroinvertebrate community metrics and caddisfly 

tissue levels. Negative correlations were also found between Cd and Zn in the water and 

sediment and the macroinvertebrate community metrics. 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 tissue 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 the 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 community extent, community composition, trophic structure, 

community dynamics, and physical structure as essential ecological attributes for assessing 

ecosystem health. The following two sections describe the effects of Pb on community 

composition, community dynamics, and trophic structure. To date, no available studies were 

located on the effects of Pb on physical structure (e.g., change in riparian tree canopy height, 

ecosystem succession). 

Ecosystems and Communities, Community Composition 

To measure community composition, an inventory of the species/taxa found in the 

ecological system must be conducted. According 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 species (Young and Sanzone, 2002). Focal or special 

species of interest can be those that play a critical 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 number of individuals of a species within 

a given area or community) 

o Reduced species richness (i.e., the number of different species present in 

a community) 

o Reduced species diversity (i.e., a measure 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 general 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, estuaries, wetlands, and species interaction. 

Aquatic Microcosm Studies 

The examination of simulated aquatic ecosystems (i.e., microcosms) provides limited 

information on the effects of pollutants on natural systems. Microcosm studies typically focus 

on only a few aspects of the natural system and do not incorporate all of the ecological, 

chemical, or biological interactions. Nevertheless, a few microcosm studies have been 

conducted that indicate potential effects 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 simulated aquatic ecosystems and found a reduction in 

the abundance and composition of protozoan species 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 exposure reduced protozoan abundance, biomass, and 
 AX7-200 Table AX7-2.5.2. Essential Ecological Attributes for Natural Aquatic Ecosystems Affected b

Category Species Condition Measures Exposure Medium Location Exposure Concentrations Oth

Met

Pres

Biotic Condition Ecosystems and Communities- Community Composition Protozoan community Reduced species abundance and diversity Marine water Laboratory microcosm 0.02– 0.05 mg/L N 

Protozoan community Reduced species abundance Freshwater water Laboratory microcosm 0.05–1 mg/L N 

Protist community Reduced species abundance and diversity Marine water Laboratory microcosm 1 mg/L N 

Meiofauna community Reduced abundance Marine sediment Laboratory microcosm 177 mg/kg dw Y 

Algal community Increased respiration Freshwater Domestic water stabilization pond 25–80 mg/L ? 

Algal community Decreased primary productivity Freshwater Sharana Basaveshwara Tank, India 6–32 mg/L ? 

Meiobenthic community (primarily nematodes) Reduced species abundance No effect on abundance Marine sediment Laboratory microcosm 1343 mg/kg dw 1580 mg/kg dw N 
 AX7-201 Table AX7-2.5.2 (cont’d). Essential Ecological Attributes for Natural Aquatic Ecosystems Affect

Category Species Condition Measures Exposure Medium Location Exposure Concentrations Oth

Met

Pres

Macroinvertebrate community Lower total abundance, decreased taxa, and EPT richness, larger percentage of tolerant species of benthic macroinvertebrates. Freshwater and sediment Mining sites in the Upper Colorado Basin <0.001–0.02 mg/L 145-850 mg/kg dw (<63 μM fraction) Y Macroinvertebrate community Negatively correlated with species richness and diversity indices Estuary sediment Douro Estuary, Portugal 0.25–192 mg/kg dw Y 

Macroinvertebrate community Reduced species abundance Freshwater sediment River Ill and tributaries, France 1–16 mg/kg dw Y 

Fish, crustacean and macroinvertebrate community Correlation with changes in species abundance and distribution Marine Sediment Spencer Gulf, South Australia 156–5270 mg/kg dw Y 

Chironomid community Reduced chironomid richness Whole organism residue New Brunswick, Canada 40.3–1,387 mg/kg dw (periphyton) 1.6–131 mg/kg dw (chironomid tissue) Y 

Macroinvertebrate community Lead in tissues negatively correlated with taxa richness, EPT richness, chironomid richness, and species density. Whole organism residue Clark Fork River, MT 32.2–67.1 mg/kg dw Y 

Macroinvertebrate community Lead in tissues negatively correlated with EPT richness and abundance. Biofilm residues Boulder River, MT 32–1540 mg/kg dw Y 
 AX7-202 Table AX7-2.5.2 (cont’d). Essential Ecological Attributes for Natural Aquatic Ecosystems Affect

Category Species Condition Measures Exposure Medium Location Exposure Concentrations Oth

Meta

Prese

 Macroinvertebrate Community Lead in tissues and sediment not correlated to diversity and richness Sediment and whole organism residue Aquashicola Creek tributaries, Palmerton, PA 7.5–59.5 mg/kg dw (sediment) 0.25-6.03 mg/kg dw (macroinvertebrates) Y 

Fish Community Lead in tissues and sediment not correlated to diversity and richness Sediment and whole organism residue Aquashicola Creek tributaries, Palmerton, PA 7.5–59.5 mg/kg dw (sediment) 0.1-0.86 mg/kg dw (fish) Y 

Ecosystems and Communities- Community Dynamics and Trophic Structure Snails and tadpoles Lead affected predator- prey interactions Sediment Outdoor mini- ecosystems Not cited Y 

Snails and caddisflies No avoidance of predator by snail. Caddisfly did respond to predator Water Field microcosm for snail; in-stream disturbance for caddisfly 27.7–277.6 mg/kg dw (snail tissue) 223–13,507 mg/kg dw (caddisfly tissue) Y 

Fathead minnow Feeding behavior altered Water Laboratory microcosm 0.5–1.0 mg/L N 

American toad No avoidance of lead Water Laboratory microcosm 0.5–1.0 mg/L N 

Mummichog Feeding behavior altered and predator avoidance affected Water Laboratory 0.3–1.0 mg/L N 
 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 effects of Pb on an estuarine meiobenthic 

community (mainly nematodes) in a microcosm setting using sediment samples collected 

offshore from England. A multivariate analysis of similarities (ANOSIM) test with square roottransformed data was used to evaluate differences 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. None 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 study, the effects of a mixture of metals (Cu, Cd, 

Pb, Hg, and Zn) on a salt marsh meiofaunal community were evaluated (Millward et al., 2001). 

After 30-days exposure, significant reductions in copepod, gastropod, and bivalve 

abundances were observed at the highest Pb exposure concentration, 177 mg/kg dw. Ostracods 

and nematodes were not affected. The authors believed that the response of the meiofauna taxa 

to metals was in part due to the various feeding strategies in that deposit feeders were most 

affected. 

Natural Aquatic Ecosystem Studies 

Lead stress in aquatic ecosystems has also been evaluated in natural communities. 

Studies examining community-scale endpoints, however, are complex, and interpretation can be 

confounded by the variability found in natural systems 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 contamination 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 presence of one or more metals (or total metals) 

and the negative effects observed. While, correlation may imply a relationship between two 
 AX7-204

variables, it does not imply causation 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. Concentrations of Pb in water (6 to 80 mg/L) were 

found to significantly reduce primary productivity and increase respiration. The authors 

suggested that increased respiration 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 macroinvertebrate community in mine-affected waters of 

Colorado. Initially, transplanted 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 μg/g dw) were higher than at reference sites (2.5 to 14.7 μg/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 reduced at mining sites. Lead levels along with 

Cd, Cu, and Zn were correlated with reduced 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 decreased in areas with 

increased Pb concentrations in the sediment. Interpretation of Pb effects was complicated by 

other non-metal stressors, namely sediment particle 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 Ill River and its tributaries in France were 

evaluated by Rosso et al. (1994). Lead in sediment (5 to 16 μg/g dw at affected sites) was 

positively correlated to the abundance of the oligochaete, Nais sp., and negatively correlated to 
 AX7-205

Tubificidae abundance. Lead was the only metal that was positively correlated to Nais species, 

while other metals were negatively correlated to Tubificidae (Rosso et al., 1994). 

The effects of metals and particle size on structuring epibenthic sea grass fauna (fish, 

mollusks, crustaceans, and polychaetes) was evaluated 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. Species 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 negative effects on species richness 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 bioaccumulation of lead in aquatic systems with 

indices of community structure and function. A focused study on changes in Chironomidae 

community composition in relation to metal mines (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 water 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) compared 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 undetectable. Chironomidae richness was reduced at the 

sites receiving mining effluent containing Pb, Cd, Cu, and Zn. 

In another study, macroinvertebrate 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 density (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 by 
 AX7-206

Poulton et al. (1995), the effects on richness and abundance of EPT taxa were compared to metal 

concentrations in tissue (i.e., biofilm and macroinvertebrates). 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 tissue 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, macroinvertebrates, 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 deposited 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 smelter. Sites were chosen, from 7.8 to 24.6 km 

from the smelter. There were no clear associations 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 that 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, predation, and succession (Young and Sanzone, 2002). 

Measures of biotic interactions (e.g., levels of seed dispersal, 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 species/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. Therefore, 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., reduced avoidance of predators) 

• Prey consumption rate (e.g., increase 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 sediment from a metals-contaminated Superfund 

site (i.e., Pb, Zn, Cd). Previous investigations of aquatic invertebrates and 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 presence of metal-contaminated sediment: 

Snails and tadpoles have similar dietary behaviors. Thus, when placed in the same habitat 

they will compete for the same food items and negatively 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 ingestion, while snails increased ingestion. Thus, 

snails were exposed to greater quantities of metals in sediment. 

In an uncontaminated environment, snail recruitment (i.e., reproduction) was reduced in 

the presence of tadpoles. The addition of tadpoles 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 effects 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 indirect in that the tadpoles hid in the algal mats 

forcing the snails to feed primarily on the benthic algae that grew on the sediment with high 
 AX7-208

metal levels. Furthermore, although not significant, 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 assessed. Snails (control and lead-exposed) were 

stimulated with a predator indicator (i.e., crushed snails and an extract of crushed snail). Control 

snails changed behaviors in the presence of the predator indicator, while exposed snails did not 

alter their behavior. The authors 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 from 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 implying previous exposure to these metals. 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 the affected lakes more prone to predation. 

The caddisflies were evaluated at 36 sites from six different streams. As with the snails, 

the caddisflies from the affected streams had higher cadmium, Pb and zinc tissue levels. The 

time for caddisfly larvae to respond (i.e., how long immobile) to disturbance (i.e., lifted from 

water for 3 seconds and moved to a new location) was evaluated. There was no correlation 

between tissue metal level and any response variable (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 testing period (4 weeks total exposure). Feeding 

behavior was evaluated by presenting two 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, exposure to Pb delayed the altering 

of prey size choices to less energetically costly prey. 
 AX7-209

Lefcort et al. (1998) exposed 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 regardless 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 effect of Pb exposure 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 affected 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 the ability of prey to 

avoid predators. The effect of Pb on these ecological 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 atmospheric Pb on ecological 

condition remains to be defined. The evaluation of point source Pb within the EPA Ecological 

Condition Framework has been examined primarily in relation to biotic conditions. The 

available literature focuses on studies describing the effects of Pb in natural aquatic ecosystems 

with regard to community composition and species interactions. The effects of Pb on the biotic 

condition of natural aquatic systems can be summarized 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 defined 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 flow and nutrient cycling and benthic community 

structure. In field studies, Pb contamination has been shown to significantly alter the aquatic 
 AX7-210

environment through bioaccumulation and alterations of community structure and function. 

Exposure to Pb in laboratory studies and simulated ecosystems may alter species competitive 

behaviors, predator-prey interactions, and contaminant avoidance behaviors. Alteration of these 

interactions may have negative effects on species 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 challenging given that observed 

effects may be due to cumulative toxicity from multiple stressors. 
 AX7-211

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