Handbook of Chemical Risk Assessment

HANDBOOK OF CHEMICAL RISK ASSESSMENT Health Hazards to Humans, Plants, and Animals VOLUME 1 Metals © 2000 by CRC Press...

Author: Ronald Eisler

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HANDBOOK OF

CHEMICAL RISK ASSESSMENT Health Hazards to Humans, Plants, and Animals VOLUME 1 Metals

© 2000 by CRC Press LLC

HANDBOOK OF

CHEMICAL RISK ASSESSMENT Health Hazards to Humans, Plants, and Animals VOLUME 1 Metals

Ronald Eisler, Ph.D. U.S. Geological Survey Patuxent Wildlife Research Center Laurel, Maryland

LEWIS PUBLISHERS Boca Raton London New York Washington, D.C.

HANDBOOK OF

CHEMICAL RISK ASSESSMENT Health Hazards to Humans, Plants, and Animals VOLUME 2 Organics


HANDBOOK OF

CHEMICAL RISK ASSESSMENT Health Hazards to Humans, Plants, and Animals VOLUME 2 Organics



HANDBOOK OF

CHEMICAL RISK ASSESSMENT Health Hazards to Humans, Plants, and Animals VOLUME 3 Metalloids, Radiation, Cumulative Index to Chemicals and Species


HANDBOOK OF

CHEMICAL RISK ASSESSMENT Health Hazards to Humans, Plants, and Animals VOLUME 3 Metalloids, Radiation, Cumulative Index to Chemicals and Species



Library of Congress Cataloging-in-Publication Data

Eisler, Ronald, 1932– Handbook of chemical risk assessment: health hazards to humans, plants, and animals / by Ronald Eisler p. cm. Includes bibliographical references and index. Contents: v. 1. Metals — v. 2. Organics — v. 3. Metalloids, radiation, cumulative index to chemicals and species. ISBN 1-56670-503-7 (v. 1: alk. paper) — ISBN 1-56670-504-5 (v. 2: alk. paper) — ISBN 1-56670-505-3 (v. 3: alk. paper) — ISBN 1-56670-506-1 (set: alk. paper) 1. Pollutants — Handbooks, manuals, etc. 2. Hazardous substances — Risk assessment — Handbooks, manuals, etc. 3. Environmental chemistry — Handbooks, manuals, etc. 4. Environmental risk assessment — Handbooks, manuals, etc. I. Title. TD196.C45 E38 2000 363.17′9 — dc21

99-053394 CIP

This work is a reprint of a publication of the United States Government. Reasonable efforts have been made to publish reliable data and information, however, neither the United States Government nor the publisher makes any warranty, express or implied, or assumes any legal liability or responsibility for the accuracy, completeness, or usefulness of any information, apparatus, product, or process disclosed, or represents that its use would not infringe privately owned rights. Reference herein to any specific commercial products, process, or service by trade name, trademark, manufacturer, or otherwise does not necessarily constitute or imply its endorsement, recommendation, or favoring by the United States Government or the publisher. The views and opinions expressed herein do not necessarily state or reflect those of the United States Government or the publisher and shall not be used for advertising or product endorsement purposes. Permission to reproduce wholly or in part for purposes other than personal or research and teaching must be requested in writing. Direct all inquiries to CRC Press LLC, 2000 N. W. Corporate Blvd., Boca Raton, Florida 33431.

First CRC Press LLC Printing, 2000 Lewis Publishers is an imprint of CRC Press LLC No claim to original U.S. Government works International Standard Book Number 1-56670-505-3 (v. 3 of 3-v. set) 1-56670-506-1 (set) Library of Congress Card Number 99-053394 Printed in the United States of America 1 2 3 4 5 6 7 8 9 0 Printed on acid-free paper

About the Author Ronald Eisler received his A.B. degree from New York University in biology and chemistry, and M.S. and Ph.D. degrees from the University of Washington in aquatic sciences and radioecology, respectively. He attended the University of Miami Marine Laboratory prior to military service. He is a member of the American Chemical Society, the American Fisheries Society, the Marine Biological Association of the United Kingdom, and the Society for Environmental Geochemistry and Health. Since 1955 he has authored more than 125 technical articles and books on contaminant hazards to living organisms. From 1984 to the present, Dr. Eisler has been a senior research biologist at the U.S. Geological Survey, Patuxent Wildlife Research Center, located in Laurel, Maryland. Prior to 1984, he held the following positions, in order: bioscience advisor, U.S. Fish and Wildlife Service, Washington, D.C.; research aquatic toxicologist, U.S. Environmental Protection Agency, Narrangansett, Rhode Island; fishery research biologist, U.S. Fish and Wildlife Service, Highlands, New Jersey; radiochemist, University of Washington Laboratory of Radiation Ecology, Seattle, Washington; aquatic biologist, New York State Department of Environmental Conservation, Raybrook, New York; biochemist, U.S. Army Medical Nutrition Laboratory, Medical Service Corps, Denver, Colorado. He has held several adjunct professor appointments and was especially active at the Graduate School of Oceanography, University of Rhode Island; the Department of Oceanography and Marine Biology, Hebrew University of Jerusalem; and the Department of Biology, American University, Washington, D.C. Dr. Eisler is currently on the editorial board of Marine Ecology Progress Series and other technical journals.


Contents Chapter 1 Cadmium 1.1 Introduction 1.2 Environmental Chemistry 1.3 Concentrations in Field Collections 1.4 Lethal Effects 1.5 Sublethal Effects 1.6 Bioaccumulation 1.7 Teratogenesis, Mutagenesis, and Carcinogenesis 1.8 Recommendations 1.9 Summary 1.10 Literature Cited Chapter 2 Chromium 2.1 Introduction 2.2 Environmental Chemistry 2.3 Concentrations in Field Collections 2.4 Beneficial and Protective Properties 2.5 Lethal Effects 2.5.1 General 2.5.2 Aquatic Organisms 2.5.3 Terrestrial Invertebrates 2.5.4 Mammals and Birds 2.6 Sublethal Effects 2.6.1 General 2.6.2 Aquatic Organisms: Freshwater 2.6.3 Aquatic Organisms: Marine 2.6.4 Birds 2.6.5 Mammals 2.7 Field Investigations 2.8 Recommendations 2.9 Summary 2.10 Literature Cited Chapter 3 Copper 3.1 Introduction 3.2 Sources and Uses 3.2.1 General 3.2.2 Sources 3.2.3 Uses 3.3 Chemical and Biochemical Properties 3.3.1 General 3.3.2 Chemical Properties 3.3.3 Metabolism 3.3.4 Interactions 3.4 Carcinogenicity, Mutagenicity, and Teratogenicity 3.4.1 General 3.4.2 Carcinogenicity


3.5

3.6

3.7

3.8 3.9 3.10

3.4.3 Mutagenicity 3.4.4 Teratogenicity Concentrations in Field Collections 3.5.1 General 3.5.2 Abiotic Materials 3.5.3 Terrestrial Plants and Invertebrates 3.5.4 Aquatic Organisms 3.5.5 Amphibians and Reptiles 3.5.6 Birds 3.5.7 Mammals Copper Deficiency Effects 3.6.1 General 3.6.2 Terrestrial Plants and Invertebrates 3.6.3 Aquatic Organisms 3.6.4 Birds and Mammals Lethal and Sublethal Effects 3.7.1 General 3.7.2 Terrestrial Plants and Invertebrates 3.7.3 Aquatic Organisms 3.7.4 Birds 3.7.5 Mammals Proposed Criteria and Recommendations Summary Literature Cited

Chapter 4 Lead 4.1 Introduction 4.2 Sources and Uses 4.3 Chemical Properties 4.4 Mode of Action 4.5 Concentrations in Field Collections 4.5.1 General 4.5.2 Nonbiological Samples 4.5.3 Fungi, Mosses, Lichens 4.5.4 Terrestrial Plants 4.5.5 Terrestrial Invertebrates 4.5.6 Aquatic Biota 4.5.7 Amphibians and Reptiles 4.5.8 Birds 4.5.9 Mammals 4.6 Lethal and Sublethal Effects 4.6.1 General 4.6.2 Terrestrial Plants and Invertebrates 4.6.3 Aquatic Biota 4.6.4 Amphibians and Reptiles 4.6.5 Birds 4.6.6 Mammals 4.7 Recommendations 4.8 Summary 4.9 Literature Cited


Chapter 5 Mercury 5.1 Introduction 5.2 Sources of Environmental Mercury 5.3 Chemical and Biochemical Properties 5.4 Mercury in Minamata, Japan 5.5 Concentrations in Field Collections 5.5.1 General 5.5.2 Nonbiological 5.5.3 Biological 5.6 Lethal Effects 5.6.1 General 5.6.2 Aquatic Organisms 5.6.3 Birds 5.6.4 Mammals 5.6.5 Other Groups 5.7 Sublethal Effects 5.7.1 General 5.7.2 Carcinogenicity, Genotoxicity, and Teratogenicity 5.7.3 Aquatic Organisms 5.7.4 Birds 5.7.5 Mammals 5.7.6 Other Groups 5.8 Recommendations 5.9 Summary 5.10 Literature Cited Chapter 6 Nickel 6.1 Introduction 6.2 Sources and Uses 6.2.1 General 6.2.2 Source 6.2.3 Uses 6.3 Chemical and Biological Properties 6.3.1 General 6.3.2 Physical and Chemical Properties 6.3.3 Metabolism 6.3.4 Interactions 6.4 Carcinogenicity, Mutagenicity, and Teratogenicity 6.4.1 General 6.4.2 Carcinogenicity 6.4.3 Mutagenicity 6.4.4 Teratogenicity 6.5 Concentrations in Field Collections 6.5.1 General 6.5.2 Abiotic Materials 6.5.3 Terrestrial Plants and Invertebrates 6.5.4 Aquatic Organisms 6.5.5 Amphibians 6.5.6 Birds


6.6

6.7

6.8 6.9 6.10

6.5.7 Mammals 6.5.8 Integrated Studies Nickel Deficiency Effects 6.6.1 General 6.6.2 Bacteria and Plants 6.6.3 Birds 6.6.4 Mammals Lethal and Sublethal Effects 6.7.1 General 6.7.2 Terrestrial Plants and Invertebrates 6.7.3 Aquatic Organisms 6.7.4 Birds 6.7.5 Mammals Proposed Criteria and Recommendations Summary Literature Cited

Chapter 7 Silver 7.1 Introduction 7.2 Sources and Uses 7.2.1 General 7.2.2 Sources 7.2.3 Uses 7.3 Chemistry and Metabolism 7.3.1 General 7.3.2 Physical and Chemical Properties 7.3.3 Metabolism 7.4 Concentrations in Field Collections 7.4.1 General 7.4.2 Nonbiological Materials 7.4.3 Plants and Animals 7.5 Lethal and Sublethal Effects 7.5.1 General 7.5.2 Terrestrial Plants 7.5.3 Aquatic Organisms 7.5.4 Birds and Mammals 7.6 Recommendations 7.7 Summary 7.8 Literature Cited Chapter 8 Tin 8.1 Introduction 8.2 Chemical and Biochemical Properties 8.2.1 General 8.2.2 Inorganic Tin 8.2.3 Organotins 8.3 Sources and Uses 8.4 Concentrations in Field Collections 8.4.1 General


8.5

8.6 8.7 8.8

8.4.2 Nonbiological Samples 8.4.3 Biological Sample Effects 8.5.1 General 8.5.2 Aquatic Organisms 8.5.3 Birds 8.5.4 Mammals 8.5.5 Terrestrial Invertebrates Recommendations Summary Literature Cited

Chapter 9 Zinc 9.1 Introduction 9.2 Sources and Uses 9.3 Chemical and Biochemical Properties 9.3.1 General 9.3.2 Chemical Properties 9.3.3 Metabolism 9.3.4 Interactions 9.4 Carcinogenicity, Mutagenicity, and Teratogenicity 9.4.1 General 9.4.2 Carcinogenicity 9.4.3 Mutagenicity 9.4.4 Teratogenicity 9.5 Concentrations in Field Collections 9.5.1 General 9.5.2 Nonbiological 9.5.3 Terrestrial Plants and Invertebrates 9.5.4 Aquatic Organisms 9.5.5 Birds 9.5.6 Mammals 9.6 Zinc Deficiency Effects 9.6.1 General 9.6.2 Terrestrial Plants 9.6.3 Aquatic Organisms 9.6.4 Birds 9.6.5 Mammals 9.7 Lethal and Sublethal Effects 9.7.1 General 9.7.2 Terrestrial Plants and Invertebrates 9.7.3 Aquatic Organisms 9.7.4 Birds 9.7.5 Mammals 9.8 Recommendations 9.9 Summary 9.10 Literature Cited


Contents Chapter 10 Acrolein 10.1 Introduction 10.2 Sources and Uses 10.2.1 General 10.2.2 Sources 10.2.3 Uses 10.3 Environmental Chemistry 10.3.1 General 10.3.2 Chemical Properties 10.3.3 Persistence 10.3.4 Metabolism 10.4 Lethal and Sublethal Effects 10.4.1 General 10.4.2 Terrestrial Plants and Invertebrates 10.4.3 Aquatic Organisms 10.4.4 Birds 10.4.5 Mammals 10.5 Recommendations 10.6 Summary 10.7 Literature Cited Chapter 11 Atrazine 11.1 Introduction 11.2 Environmental Chemistry 11.3 Concentrations in Field Collections 11.4 Effects 11.4.1 General 11.4.2 Terrestrial Plants and Invertebrates 11.4.3 Aquatic Plants 11.4.4 Aquatic Animals 11.4.5 Birds 11.4.6 Mammals 11.5 Recommendations 11.6 Summary 11.7 Literature Cited Chapter 12 Carbofuran 12.1 Introduction 12.2 Chemical Properties and Persistence 12.3 Lethal Effects 12.3.1 General 12.3.2 Aquatic Organisms 12.3.3 Birds and Mammals 12.3.4 Terrestrial Invertebrates 12.3.5 Plants 12.4 Sublethal Effects


12.4.1 General 12.4.2 Aquatic Organisms 12.4.3 Birds 12.4.4 Mammals 12.4.5 Terrestrial Invertebrates 12.5 Recommendations 12.6 Summary 12.7 Literature Cited Chapter 13 Chlordane 13.1 Introduction 13.2 Chemical and Biochemical Properties 13.3 Uses 13.4 Concentrations in Field Collections 13.4.1 General 13.4.2 Nonbiological Samples 13.4.3 Terrestrial Crops 13.4.4 Aquatic Invertebrates 13.4.5 Fishes 13.4.6 Amphibians and Reptiles 13.4.7 Birds 13.4.8 Mammals 13.5 Lethal and Sublethal Effects 13.5.1 General 13.5.2 Terrestrial Invertebrates 13.5.3 Aquatic Organisms 13.5.4 Amphibians and Reptiles 13.5.5 Birds 13.5.6 Mammals 13.6 Recommendations 13.7 Summary 13.8 Literature Cited Chapter 14 Chlorpyrifos 14.1 Introduction 14.2 Environmental Chemistry 14.3 Laboratory Investigations 14.3.1 Aquatic Organisms 14.3.2 Birds and Mammals 14.4 Field Investigations 14.5 Recommendations 14.6 Summary 14.7 Literature Cited Chapter 15 Cyanide 15.1 Introduction 15.2 Chemical Properties 15.3 Mode of Action


15.4 15.5 15.6 15.7 15.8 15.9

Clinical Features Antidotes Sources and Uses Concentrations in Field Collections Persistence in Water, Soil, and Air Lethal and Sublethal Effects 15.9.1 Terrestrial Flora and Invertebrates 15.9.2 Aquatic Organisms 15.9.3 Birds 15.9.4 Mammals 15.10 Recommendations 15.11 Summary 15.12 Literature Cited Chapter 16 Diazinon 16.1 Introduction 16.2 Environmental Chemistry 16.3 Lethal Effects 16.3.1 General 16.3.2 Aquatic Organisms 16.3.3 Birds 16.3.4 Mammals 16.3.5 Terrestrial Invertebrates 16.4 Sublethal Effects 16.4.1 General 16.4.2 Aquatic Organisms 16.4.3 Birds 16.4.4 Mammals 16.4.5 Terrestrial Invertebrates 16.5 Recommendations 16.6 Summary 16.7 Literature Cited Chapter 17 Diflubenzuron 17.1 Introduction 17.2 Environmental Chemistry 17.2.1 General 17.2.2 Chemical and Biochemical Properties 17.2.3 Persistence in Soil and Water 17.3 Uses 17.4 Lethal and Sublethal Effects 17.4.1 General 17.4.2 Terrestrial Plants 17.4.3 Terrestrial Invertebrates 17.4.4 Aquatic Organisms: Laboratory Studies 17.4.5 Aquatic Organisms: Field Studies 17.4.6 Birds 17.4.7 Mammals 17.5 Recommendations


17.6 Summary 17.7 Literature Cited Chapter 18 Dioxins 18.1 Introduction 18.2 Environmental Chemistry 18.3 Concentrations in Field Collections 18.4 Lethal and Sublethal Effects 18.4.1 General 18.4.2 Terrestrial Plants and Invertebrates 18.4.3 Aquatic Organisms 18.4.4 Birds 18.4.5 Mammals 18.5 Recommendations 18.6 Summary 18.7 Literature Cited Chapter 19 Famphur 19.1 Introduction 19.2 Uses 19.3 Chemistry and Metabolism 19.4 Lethal and Sublethal Effects 19.4.1 General 19.4.2 Terrestrial Invertebrates 19.4.3 Aquatic Organisms 19.4.4 Birdst 19.4.5 Mammals 19.5 Recommendations 19.6 Summary 19.7 Literature Cited Chapter 20 Fenvalerate 20.1 Introduction 20.2 Environmental Chemistry 20.2.1 General 20.2.2 Chemical Properties 20.2.3 Uses 20.2.4 Persistence 20.3 Mode of Action 20.3.1 General 20.3.2 Types of Pyrethroids 20.3.3 Sodium Gating Kinetics 20.3.4 Metabolism 20.3.5 Mutagenicity, Teratogenicity, and Carcinogenicity 20.4 Effects 20.4.1 General 20.4.2 Terrestrial Plants and Invertebrates 20.4.3 Aquatic Organisms


20.4.4 Birds 20.4.5 Mammals 20.5 Recommendations 20.6 Summary 20.7 Literature Cited Chapter 21 Mirex 21.1 Introduction 21.2 Chemical Properties 21.3 Lethal Effects 21.3.1 Aquatic Organisms 21.3.2 Birds and Mammals 21.4 Sublethal Effects 21.4.1 Aquatic Organisms 21.4.2 Birds 21.4.3 Mammals 21.5 Bioaccumulation 21.5.1 Aquatic Organisms 21.5.2 Birds and Mammals 21.6 Mirex in the Southeastern United States 21.7 Mirex in the Great Lakes 21.8 Mirex in Other Geographic Areas 21.9 Recommendations 21.10 Summary 21.11 Literature Cited Chapter 22 Paraquat 22.1 Introduction 22.2 Uses 22.3 Concentrations in Field Collections 22.4 Environmental Chemistry 22.4.1 General 22.4.2 Chemical Properties 22.4.3 Mode of Action 22.4.4 Fate in Soils and Water 22.5 Lethal and Sublethal Effects 22.5.1 General 22.5.2 Terrestrial Plants and Invertebrates 22.5.3 Aquatic Organisms 22.5.4 Birds 22.5.5 Mammals 22.6 Recommendations 22.7 Summary 22.8 Literature Cited Chapter 23 Pentachlorophenol 23.1 Introduction 23.2 Environmental Chemistry


23.3

23.4

23.5 23.6 23.7

23.2.1 General 23.2.2 Sources and Uses 23.2.3 Properties 23.2.4 Fate Concentrations in Field Collections 23.3.1 General 23.3.2 Biological and Nonbiological Samples Effects 23.4.1 General 23.4.2 Terrestrial Plants and Invertebrates 23.4.3 Aquatic Biota 23.4.4 Birds 23.4.5 Mammals Recommendations Summary Literature Cited

Chapter 24 Polychlorinated Biphenyls 24.1 Introduction 24.2 Sources and Uses 24.3 Chemical and Biochemical Properties 24.3.1 General 24.3.2 Physical Properties 24.3.3 Toxic Equivalency Factors 24.3.4 Structure–Function Relations 24.3.5 Quantitation 24.4 Concentrations in Field Collections 24.4.1 General 24.4.2 Nonbiological Materials 24.4.3 Marine Mammals 24.4.4 Other Aquatic Organisms 24.4.5 Reptiles 24.4.6 Birds 24.4.7 Terrestrial Mammals 24.5 Lethal and Sublethal Effects 24.5.1 General 24.5.2 Aquatic Organisms 24.5.3 Birds 24.5.4 Mammals 24.6 Recommendations 24.7 Summary 24.8 Literature Cited Chapter 25 Polycyclic Aromatic Hydrocarbons 25.1 Introduction 25.2 Environmental Chemistry, Sources, and Fate 25.2.1 Properties 25.2.2 Sources 25.2.3 Fate


25.3 Concentrations in Field Collections 25.3.1 General 25.3.2 Nonbiological Samples 25.3.3 Biological Samples 25.4 Lethal and Sublethal Effects 25.4.1 General 25.4.2 Fungi 25.4.3 Terrestrial Plants 25.4.4 Aquatic Biota 25.4.5 Amphibians and Reptiles 25.4.6 Birds 25.4.7 Mammals 25.5 Recommendations 25.6 Summary 25.7 Literature Cited Chapter 26 Sodium Monofluoroacetate 26.1 Introduction 26.2 Uses 26.2.1 Domestic Use 26.2.2 Nondomestic Use 26.3 Environmental Chemistry 26.3.1 General 26.3.2 Chemical Properties 26.3.3 Persistence 26.3.4 Metabolism 26.3.5 Antidotes 26.4 Lethal and Sublethal Effects 26.4.1 General 26.4.2 Terrestrial Plants and Invertebrates 26.4.3 Aquatic Organisms 26.4.4 Amphibians and Reptiles 26.4.5 Birds 26.4.6 Mammals 26.5 Recommendations 26.6 Summary 26.7 Literature Cited Chapter 27 Toxaphene 27.1 Introduction 27.2 Environmental Chemistry 27.3 Concentrations in Field Populations 27.4 Lethal Effects 27.5 Sublethal Effects 27.6 Recommendations 27.7 Summary 27.8 Literature Cited


Contents Chapter 28 Arsenic 28.1 Introduction 28.2 Sources, Fate, and Uses 28.3 Chemical and Biochemical Properties 28.4 Essentiality, Synergism, and Antagonism 28.5 Concentrations in Field Collections 28.5.1 General 28.5.2 Nonbiological Samples 28.5.3 Biological Samples 28.6 Lethal and Sublethal Effects 28.6.1 General 28.6.2 Carcinogenesis, Mutagenesis, and Teratogenesis 28.6.3 Terrestrial Plants and Invertebrates 28.6.4 Aquatic Biota 28.6.5 Birds 28.6.6 Mammals 28.7 Recommendations 28.8 Summary 28.9 Literature Cited Chapter 29 Boron 29.1 Introduction 29.2 Environmental Chemistry 29.2.1 General 29.2.2 Sources and Uses 29.2.3 Chemical Properties 29.2.4 Mode of Action 29.3 Concentrations in Field Collections 29.3.1 General 29.3.2 Nonbiological Materials 29.3.3 Plants and Animals 29.4 Effects 29.4.1 General 29.4.2 Terrestrial Plants 29.4.3 Terrestrial Invertebrates 29.4.4 Aquatic Organisms 29.4.5 Birds 29.4.6 Mammals 29.5 Recommendations 29.6 Summary 29.7 Literature Cited Chapter 30 Molybdenum 30.1 Introduction 30.2 Environmental Chemistry


30.3

30.4

30.5 30.6 30.7

30.2.1 General 30.2.2 Sources and Uses 30.2.3 Chemical Properties 30.2.4 Mode of Action Concentrations in Field Collections 30.3.1 General 30.3.2 Nonbiological Samples 30.3.3 Biological Samples Effects 30.4.1 General 30.4.2 Terrestrial Plants 30.4.3 Terrestrial Invertebrates 30.4.4 Aquatic Organisms 30.4.5 Birds 30.4.6 Mammals Recommendations Summary Literature Cited

Chapter 31 Selenium 31.1 Introduction 31.2 Environmental Chemistry 31.3 Concentrations in Field Collections 31.4 Deficiency and Protective Effects 31.5 Lethal Effects 31.5.1 Aquatic Organisms 31.5.2 Mammals and Birds 31.6 Sublethal and Latent Effects 31.6.1 Aquatic Organisms 31.6.2 Terrestrial Invertebrates 31.6.3 Birds 31.6.4 Mammals 31.7 Recommendations 31.8 Summary 31.9 Literature Cited Chapter 32 Radiation 32.1 Introduction 32.2 Physical Properties of Radiation 32.2.1 General 32.2.2 Electromagnetic Spectrum 32.2.3 Radionuclides 32.2.4 Linear Energy Transfer 32.2.5 New Units of Measurement 32.3 Sources and Uses 32.3.1 General 32.3.2 Natural Radioactivity 32.3.3 Anthropogenic Radioactivity 32.3.4 Dispersion


32.4 Radionuclide Concentrations in Field Collections 32.4.1 General 32.4.2 Abiotic Materials 32.4.3 Aquatic Ecosystems 32.4.4 Birds 32.4.5 Mammals 32.5 Case Histories 32.5.1 Pacific Proving Grounds 32.5.2 Chernobyl 32.6 Effects: Nonionizing Radiations 32.7 Effects: Ionizing Radiations 32.7.1 General 32.7.2 Terrestrial Plants and Invertebrates 32.7.3 Aquatic Organisms 32.7.4 Amphibians and Reptiles 32.7.5 Birds 32.7.6 Mammals 32.8 Proposed Criteria and Recommendations 32.9 Summary 32.10 Literature Cited 32.11 Glossary Chapter 33 Cumulative Index to Chemicals and Species 33.1 Introduction 33.2 Index to Chemicals 33.3 Index to Species 33.4 Summary 33.5 Literature Cited


LIST OF TABLES 1.1 Cadmium burdens and residence times in the principal global reservoirs 1.2 Cadmium concentrations in field collections of selected species of plants and animals 1.3 Lethal concentrations (LC) of cadmium to freshwater biota during various exposure intervals 1.4 Maximum allowable toxicant concentrations (MATC) of cadmium to sensitive species of freshwater teleosts 1.5 Sublethal effects of cadmium to selected species of aquatic biota 1.6 Cadmium residues, in mg/kg dry weight (ppm) in soil, flora, and fauna collected near two zinc smelters in Palmerton, Pennsylvania 1.7 Bioconcentration of cadmium from ambient medium by selected species of aquatic biota 1.8 Proposed cadmium criteria for the protection of living resources and human health 2.1 Chromium concentrations in abiotic materials 2.2 Concentrations of total chromium and Cr+6 in air, water, soil, and sludge near industrial sites and sewage outfalls in the United States 2.3 Chromium concentrations in field collections of selected species of marine, freshwater, and terrestrial plants and animals 2.4 Acute toxicities of hexavalent and trivalent chromium to aquatic life 2.5 Maximum acceptable toxicant concentration (MATC) values for hexavalent and trivalent chromium to aquatic life based on life cycle or partial life cycle exposures 2.6 Proposed chromium criteria for the protection of human health and natural resources 3.1 Some properties of copper and copper sulfate 3.2 Copper concentrations in selected abiotic materials 3.3 Copper concentrations (milligrams of copper per kilogram fresh weight [FW], dry weight [DW], or ash weight [AW]) in field collections of representative plants and animals 3.4 Effects of copper on representative terrestrial plants and invertebrates 3.5 Effects of copper on representative aquatic plants and animals 3.6 Effects of copper on selected birds . 3.7 Effects of copper on selected mammals 3.8 Proposed copper criteria for the protection of natural resources and human health 4.1 4.2 4.3 4.4 4.5 4.6 4.7 4.8 4.9

World lead production, consumption, and principal end uses Use patterns for lead in selected countries Amounts of lead in global reservoirs Lead concentrations in selected nonbiological materials Lead concentrations in field collections of representative species of plants and animals Lethal and sublethal effects of lead to selected species of aquatic organisms Lethal and sublethal effects of lead to selected species of birds Lethal and sublethal effects of lead to selected species of mammals Proposed lead criteria for the protection of natural resources and human health

5.1 Industrial and other uses of mercury in the United States in 1959 and in 1969 5.2 Amount of mercury in some global reservoirs and residence time 5.3 Some properties of mercury and its compounds


5.4 Mercury concentrations in selected biological and nonbiological materials collected from Minamata Bay, Japan, and environs 5.5 Mercury concentrations in selected abiotic materials 5.6 Mercury concentrations in field collections of selected species of plants and animals 5.7 Lethality of inorganic and organic mercury compounds to selected species of aquatic organisms 5.8 Lethality to birds of mercury administered by oral, dietary, or other routes 5.9 Lethality of organomercury compounds to selected mammals 5.10 Sublethal effects of organomercury compounds administered to selected species of mammals 5.11 Proposed mercury criteria for protection of various resources and human health 6.1 Nickel chronology 6.2 World mine production of nickel 6.3 Nickel consumption in the United States by intermediate product and end-use industry in 1985 6.4 Inventory of nickel in various global environmental compartments 6.5 Nickel concentrations in selected abiotic materials 6.6 Nickel concentrations (milligrams of nickel per kilogram fresh weight [FW] or dry weight [DW]) in field collections of representative plants and animals 6.7 Nickel effects on selected aquatic plants and animals 6.8 Nickel effects on birds 6.9 Nickel effects on selected mammals 6.10 Proposed nickel criteria for protection of natural resources and human health 7.1 7.2 7.3 7.4 7.5

Estimated release of silver to the environment in the United States in 1978 Industrial domestic use of silver during 1966–72 Some properties of silver and silver nitrate Silver concentrations in representative nonbiological materials Silver concentrations (milligrams of silver per kilogram fresh weight [FW], dry weight [DW], or ash weight [AW]) in field collections of selected plants and animals 7.6 Effects of silver on representative aquatic plants and animals 7.7 Effects of silver on selected mammals 7.8 Proposed silver criteria for the protection of natural resources and human health 8.1 8.2 8.3 8.4 8.5 8.6 8.7 8.8 8.9 8.10 8.11 8.12

Biological and abiotic degradation times of selected organotins Annual tin production and consumption Major uses of inorganic and organic tin compounds Possible modes of entry of organotins into air, soil, and water Total tin flux to the atmosphere and hydrosphere Tin concentrations in nonbiological materials Tin concentrations in field collections of living flora and fauna Lethal and sublethal effects of inorganic and organic tin compounds in ambient medium to selected species of aquatic organisms Toxicity of selected diorganotin and triorganotin compounds to zoeae of the marine mud crab (Rithropanopeus harrisii) exposed from hatching to age 14 days Lethal and sublethal effects of selected organotin compounds to birds Effects of tin compounds on selected species of mammals Proposed organotin criteria for protection of natural resources and human health


9.1

Estimated annual zinc inputs to United States coastal marine ecosystems; study area 2. comprised 116,000 km 9.2 Some properties of zinc, zinc chloride, and zinc sulfate 9.3 Zinc concentrations (milligrams of zinc per kilogram fresh weight [FW] or dry weight DW]) in representative nonbiological materials 9.4 Zinc concentrations (milligrams of zinc per kilogram fresh weight [FW] or dry weight [DW]) in field collections of representative plants and animals 9.5 Effects of zinc on representative terrestrial plants and invertebrates 9.6 Effects of zinc on representative aquatic plants and animals 9.7 Effects of zinc on representative birds 9.8 Effects of zinc on representative mammals 9.9 Proposed zinc criteria for the protection of natural resources and human health 10.1 Chemical and other properties of acrolein 10.2 Acrolein effects on representative aquatic organism 10.3 Acrolein effects on birds 10.4 Acrolein effects on selected mammals 10.5 Proposed acrolein criteria for the protection of living resources and human health 11.1 11.2 11.3 11.4 11.5 11.6 11.7

Chemical and other properties of atrazine Atrazine concentrations in selected watersheds Atrazine effects on selected species of terrestrial plants Atrazine effects on selected species of aquatic plants Lethal and sublethal effects of atrazine on selected species of aquatic animals Atrazine effects on selected species of birds Lethal and sublethal effects of atrazine on selected species of mammals

12.1

12.2 12.3 12.4 12.5 12.6 12.7

Carbofuran and its degradation products (in mg/kg dry weight) in corn (Zea mays) at silage stage (117 days) and at harvest (149 days) following application of carbofuran (10%) granules at 5.41 kg/ha Effect of pH, soil type, and application rate on carbofuran degradation in soils Acute toxicities of carbofuran to aquatic organisms Acute oral toxicities of carbofuran to birds and mammals Toxicity of dietary carbofuran to birds and mammals Acute aerosol inhalation toxicity of carbofuran to warm-blooded animals Rates of carbofuran metabolism by various organisms

13.1 13.2 13.3 13.4 13.5 13.6

Chlordane concentrations in selected nonbiological samples Chlordane concentrations in field collections of selected animals Chlordane effects on selected aquatic organisms Chlordane effects on selected birds Chlordane effects on selected mammals Proposed chlordane criteria for protection of natural resources and human health

14.1 14.2 14.3

Selected chlorpyrifos formulations and carriers Chemical and other properties of chlorpyrifos Acute toxicities of chlorpyrifos to selected species of aquatic invertebrates and fishes Chlorpyrifos bioconcentration factor (BCF) of selected species of fishes Chlorpyrifos lethality to selected birds and mammals via single oral dose route of administration

14.4 14.5


14.6 14.7

Dietary toxicity of chlorpyrifos to selected species of birds Chlorpyrifos effects on selected ecosystems

15.1 15.2

Some properties of potassium cyanide, hydrogen cyanide, and sodium cyanide Concentrations of cyanide in field collections of selected living resources and nonbiological materials Cyanide effects on selected species of aquatic organisms Cyanide effects on selected species of birds Cyanide effects on selected species of mammals Proposed free cyanide criteria for the protection of living resources and human health

15.3 15.4 15.5 15.6

16.1 16.2 16.3

16.8

Persistence of diazinon in plants, soil, and water Acute toxicity of diazinon to aquatic organisms Mortality of mallard embryos after immersion for 30 seconds in graded strength diazinon solutions Acute oral toxicity of diazinon to birds and mammals Toxicity of diazinon to laboratory animals via dermal, inhalation, dietary, and chronic oral routes of administration Accumulation of diazinon by aquatic organisms Lowest tested diazinon concentrations that in medium produced significant nonlethal biological effects to aquatic organisms Sublethal effects of diazinon in selected mammals

17.1 17.2 17.3 17.4 17.5 17.6 17.7

Chemical and other properties of diflubenzuron Diflubenzuron persistence in soil and water Diflubenzuron effects on selected terrestrial invertebrates Diflubenzuron effects on selected aquatic organisms: laboratory studies Diflubenzuron effects on selected aquatic organisms: field studies Diflubenzuron effects on selected birds Diflubenzuron effects on selected mammals

18.1

Levels of PCDDs in commercial chlorinated phenols, and levels of 2,3,7,8-TCDD in 2,4,5-T acid and ester formulations Chemical and physical properties of 2,3,7,8-TCDD, also known as CAS Registry No. 1746-01-6 Concentrations of PCDDs measured in selected organisms and nonbiological materials collected from various locales Effects of 2,3,7,8-TCDD and other PCDDs on representative species of freshwater organisms Acute toxicities of selected PCDD isomers to the guinea pig and the mouse Acute oral toxicities of 2,3,7,8-TCDD to mammals Proposed 2,3,7,8-TCDD criteria for the protection of natural resources and human health

16.4 16.5 16.6 16.7

18.2 18.3 18.4 18.5 18.6 18.7 19.1 19.2 19.3 19.4

Chemical and other properties of famphur Famphur effects on selected terrestrial invertebrates Famphur effects on birds Famphur effects on mammals

20.1

Chemical and other properties of fenvalerate


20.2 20.3 20.4 20.5 20.6 20.7 20.8

Fenvalerate persistence in water, sediments, and soils Fenvalerate persistence in plants and animals under field conditions Lethal and sublethal effects of fenvalerate on terrestrial invertebrates Lethal and sublethal effects of fenvalerate on aquatic organisms Lethal and sublethal effects of fenvalerate on birds Lethal and sublethal effects of fenvalerate on mammals Proposed fenvalerate criteria for the protection of natural resources and human health

21.1 21.2 21.3 21.4

Acute oral toxicity of mirex to birds and mammals Dietary toxicity of mirex to vertebrate organisms Uptake of mirex from ambient medium or diet by selected species Mirex residues in water, sediments, and fauna in a South Carolina coastal marsh 18 months after application of 4.2 g/ha Temporal persistence of residues for 1 year after applications of mirex 4X formulation to bahia grass pastures Mirex residues in fauna near Jacksonville, Florida, at various intervals posttreatment following single application of 1.12 g mirex/ha Mirex and its degradation products in herring gull eggs collected from the Great Lakes in 1977 Biomagnification of mirex in Great Lakes food chains

21.5 21.6 21.7 21.8 22.1 22.2 22.3 22.4 22.5 22.6

23.1 23.2 23.3 23.4 23.5 23.6 23.7

24.1 24.2 24.3 24.4 24.5 24.6

Paraquat concentrations in field collections of selected organisms and nonbiological materials Chemical and other properties of paraquat Effects of paraquat on selected species of aquatic plants and animals Effects of paraquat on selected species of birds Effects of paraquat on selected species of mammals Proposed paraquat criteria for the protection of natural resources and human health Chemical and other properties of pentachlorophenol Partial list of contaminants detected in technical-grade and purified pentachlorophenol Pentachlorophenol (PCP) concentrations in nonbiological and living materials Lethal and sublethal effects of pentachlorophenol (PCP) on selected species of aquatic organisms Effects of pentachlorophenol (PCP) on selected species of birds Effects of pentachlorophenol on selected mammals Proposed pentachlorophenol criteria for the protection of natural resources and human health Estimated PCB loads in the global environment Polychlorinated biphenyls: structure, retention times, response factors, and octanol/water partition coefficients (log Kow) Proposed toxicity equivalency value (TEF) relative to 2,3,7,8-TCDD of non-ortho, mono-ortho, and di-ortho planar PCBs Interactive effects of PCBs on the induction of rat (Rattus sp.) liver microsomal cytochrome P-450c Summary of PCB structure-function relations in rats (Rattus sp.) PCB congeners in Aroclor 1254 and 1260


24.7 24.8 24.9 24.10 24.11 24.12 24.13 24.14 24.15 24.16

Effect of oven drying of sediments on percent loss of selected PCB congeners present at >0.1 mg/kg fresh weight PCB concentrations in field collections of selected nonbiological materials PCB concentrations in field collections of selected aquatic organisms PCB concentrations in field collections of selected reptiles PCB concentrations in field collections of selected birds PCB concentrations in field collections of selected mammals Effects of PCBs on representative aquatic organisms Effects of PCBs on selected birds Effects of PCBs on selected mammals Proposed PCB criteria for the protection of natural resources and human health

25.1 25.2 25.3 25.4 25.5 25.6 25.7 25.8 25.9

Some physical and chemical properties of selected PAHs Major sources of PAHs in atmospheric and aquatic environments PAH concentrations in selected nonbiological materials PAH concentrations in field collections of selected biota Toxicity of selected PAHs to aquatic organisms PAH bioconcentration factor (BCF) for selected species of aquatic organisms Benzo[a]pyrene effects on selected aquatic vertebrates Some effects of PAHs on selected laboratory mammals Proposed PAH criteria for protection of human health and aquatic life

26.1 26.2 26.3 26.4

Some properties of sodium monofluoroacetate Effects of 1080 on representative amphibians and reptiles Effects of 1080 on representative birds Effects of 1080 on representative mammals

27.1

Toxaphene concentrations in field collections of selected living organisms and abiotic materials Acute toxicity of toxaphene to aquatic organisms Maximum acceptable toxicant concentration values (MATC) for toxaphene and aquatic organisms, based on exposure for the entire or most of the life cycle Acute oral toxicity of toxaphene to birds and mammals Sublethal effects of toxaphene to aquatic biota Bioconcentration factor (BCF) for toxaphene and selected species of aquatic biota Sublethal effects of toxaphene on selected mammals Proposed toxaphene criteria for protection of natural resources and human health Total arsenic concentrations in selected nonbiological materials Arsenic concentrations in field collections of selected species of plants and animals Lethal and sublethal effects of various arsenic compounds on selected species of terrestrial plants and invertebrates Lethal and sublethal effects of various arsenic compounds on selected species of aquatic biota Lethal and sublethal effects of various arsenicals on selected species of birds Lethal and sublethal effects of various arsenicals on selected species of mammals Proposed arsenic criteria for the protection of selected natural resources and human health

27.2 27.3 27.4 27.5 27.6 27.7 27.8 28.1 28.2 28.3 28.4 28.5 28.6 28.7

29.1

Environmental sources of domestic boron


29.2 29.3

Boron concentrations in selected nonbiological materials Boron concentrations in field collections of selected species of plants and animals 29.4 Boron toxicity to some terrestrial plants 29.5 Boron concentrations in soil water associated with optimal growth and plant injury 29.6 Lethal and sublethal effects of boron on terrestrial invertebrates 29.7 Lethal and sublethal effects of boron on aquatic organisms 29.8 Lethal and sublethal effects of boron on birds 29.9 Lethal and sublethal effects of boron on mammals 29.10 Proposed boron criteria for the protection of natural resources and human health 30.1 30.2 30.3 30.4 30.5

31.1 31.2 31.3 31.4 31.5

32.1 32.2 32.3 32.4 32.5 32.6 32.7 32.8 32.9 32.10 32.11 32.12 32.13 32.14 32.15

Molybdenum concentrations in selected nonbiological materials Molybdenum concentrations in field collections of selected species of animals and plants Molybdenum effects on selected species of aquatic organisms Molybdenum effects on selected species of mammals Proposed molybdenum criteria for the protection of living resources and human health Selenium concentrations in nonbiological materials Selenium concentrations in field populations of selected species of flora and fauna Toxicity of selenium salts to aquatic biota Selenium effects on birds Proposed criteria for prevention of selenium deficiency and for protection against selenosis Selected radionuclides: symbol, mass number, atomic number, half-life, and decay mode Radiation weighting factors for various types of ionizing radiations New units for use with radiation and radioactivity Annual effective dose equivalent to humans from natural sources of ionizing radiation Annual whole-body radiation doses to humans from various sources Sources and applications of atomic energy Annual effective dose equivalent from nuclear weapons testing to humans in the north temperate zone Estimated fallout of 90Sr and 137Cs over the Great Lakes, 1954–1983, in cumulative millions of Bq/km2 Fission products per kg 235U reactor charge at 100 days cooling Radioactive waste disposal at sea Theoretical peak dose, in microsieverts per year, received from plutonium and americium by three human populations Time required to transport selected radionuclides added into marine waters at surface from the upper mixed layer by biological transport Radionuclide concentrations in field collections of selected abiotic materials Radionuclide concentrations in field collections of selected living organisms Radionuclide concentrations in selected samples from the Pacific Proving Grounds


32.16 Selected fission products in the Chernobyl reactor core and their estimated escape into the environment 32.17 Regional total effective human dose-equivalent commitment from the Chernobyl accident 32.18 Radionuclide concentrations in biotic and abiotic materials from various geographic locales before or after the Chernobyl nuclear accident on April 26, 1986 32.19 Radiation effects on selected terrestrial plants 32.20 Radiation effects on selected terrestrial invertebrates 32.21 Radiation effects on selected aquatic organisms 32.22 Concentration factors for cesium-137 and strontium-90 in aquatic organisms 32.23 Approximate maximum concentration factors for selected transuranics in marine sediments, macroalgae, and fish 32.24 Maximum concentration factors reported for selected elements in marine organisms at various trophic levels 32.25 Radiation effects on selected amphibians and reptiles 32.26 Radiation effects on selected birds 32.27 Radiation effects on selected mammals 32.28 Recommended radiological criteria for the protection of human health 33.1 33.2

Chemical and trade names of substances listed in the Handbook of Chemical Risk Assessment series Common and scientific names of plants and animals listed in the Handbook of Chemical Risk Assessment series


LIST OF FIGURES 5.1 Major transformations of mercury in the environment 5.2 Mercury concentrations in livers of four species of pinniped mammals 8.1 Environmental degradation scheme for tributyltin and triphenyltin compounds 10.1 Proposed scheme for in vitro mammalian metabolism of acrolein and allyl alcohol, a precursor of acrolein 11.1 Structural formula of atrazine 12.1 Structural formula for carbofuran (2,3-dihydro-2,2-dimethyl-1,7-benzofuranyl methylcarbamate) 13.1 Chemical structure of chlordane-related compounds 14.1 Structures of chlorpyrifos and some of its metabolites 15.1 Summary of lethal and sublethal effects of free cyanide on freshwater fish 16.1 Structural formula of diazinon 17.1 Generalized degradation pattern for diflubenzuron 18.1 Upper: numbering system used for identification of individual PCDD isomers. Lower: the isomer 2,3,7,8-tetrachlorodibenzo-para-dioxin (2,3,7,8-TCDD) 19.1 Metabolic scheme for famphur in mammals 20.1 Fenvalerate and its isomers 22.1 Structural formulas of paraquat cation (upper) and paraquat dichloride salt (lower) 22.2 Proposed pathway of paraquat degradation by a bacterial isolate (upper) and by ultraviolet irradiation (lower) 23.1 Structural formula of pentachlorophenol 23.2 Some impurities found in technical-grade pentachlorophenol 24.1 Structure of biphenyl 24.2 Planar polychlorinated biphenyls and their derivatives 25.1 Nomenclature of PAHs 25.2 Ring structures of representative noncarcinogenic PAHs 25.3 Ring structures of representative tumorigenic, cocarcinogenic, and carcinogenic PAHs 25.4 The bay region dihydrodiol epoxide route of benzo[a]pyrene 32.1 The spectrum of electromagnetic waves, showing the relationship between wavelength, frequency, and energy 32.2 The principal uranium-238 decay series, indicating major decay mode and physical half-time of persistence 32.3 The three still-existing natural decay series 32.4 Natural radiations in selected radiological domains 32.5 Plutonium-239+240 in environmental samples at Thule, Greenland, between 1970 and 1984, after a military accident in 1968 32.6 Chernobyl air plume behavior and reported initial arrival times of detectable radioactivity 32.7 Acute radiation dose range fatal to 50% (30 days postexposure) of various taxonomic groups 32.8 Relation between diet, metabolism, and body weight with half-time retention of longest-lived component of cesium-137 32.9 Survival time and associated mode of death of selected mammals after whole-body doses of gamma radiation 32.10 Relation between body weight and radiation-induced LD50 (30 days postexposure) for selected mammals


Contents of Volumes 1, 2 and 3 Volume 1 Metals 1. Cadmium 2. Chromium 3. Copper 4. Lead 5. Mercury 6. Nickel 7. Silver 8. Tin 9. Zinc

Volume 2 Organics 10. Acrolein 11. Atrazine 12. Carbofuran 13. Chlordane 14. Chlorpyrifos 15. Cyanide 16. Diazinon 17. Diflubenzuron 18. Dioxins 19. Famphur 20. Fenvalerate 21. Mirex 22. Paraquat 23. Pentachlorophenol 24. Polychlorinated Biphenyls 25. Polycyclic Aromatic Hydrocarbons 26. Sodium Monofluoroacetate 27. Toxaphene

Volume 3 Metalloids, Radiation, Cumulative Index to Chemicals and Species 28. Arsenic 29. Boron 30. Molybdenum 31. Selenium 32. Radiation 33. Cumulative Index to Chemicals and Species


Preface Risk assessment is an inexact science. Successful risk assessment practitioners rely heavily on extensive and well-documented databases. In the case of chemicals entering the environment as a result of human activities, these databases generally include the chemical’s source and use; its physical, chemical, and metabolic properties; concentrations in field collections of abiotic materials and living organisms; deficiency effects, in the case of chemicals essential for life processes; lethal and sublethal effects, including effects on survival, growth, reproduction, metabolism, mutagenicity, teratogenicity, and carcinogenicity; proposed regulatory criteria for the protection of human health and sensitive natural resources; and recommendations for additional research when databases are incomplete. Of the hundreds of thousands of chemicals discharged into the environment yearly from agricultural, domestic, industrial, mining, manufacturing, municipal, and military operations, there is sufficient information on only a small number to attempt preliminary risk assessments. The chemicals selected for inclusion in this handbook series were recommended by environmental specialists of the U.S. Fish and Wildlife Service and other resource managers responsible for protecting our fish, wildlife, and biological diversity. Their choices were based on real or potential impact of each contaminant on populations of free-living natural resources — including threatened and endangered species — and on insufficient knowledge on how best to mitigate damage effects. Each chapter selectively reviews and synthesizes the technical literature on a specific priority contaminant in the environment and its effects on notably terrestrial plants and invertebrates, aquatic plants and animals, avian and mammalian wildlife, and other natural resources. Early versions of individual chapters were published between 1985 and 1999 in the Contaminant Hazard Reviews series of the U.S. Department of the Interior. This series rapidly became an important reference tool for contaminant specialists, scientists, educational institutions, state and local governments, natural resources agencies, business and industry, and the general public. More than 105,000 copies of individual reports were distributed in response to specific requests before supplies became exhausted. The current offering is updated to reflect the burgeoning literature in this subject area. I authored the original versions and the updates while stationed at the Patuxent Wildlife Research Center; however, all interpretations are my own and do not necessarily reflect those of the U.S. government. Moreover, mention of trade names or commercial products is not an endorsement or recommendation for use by the U.S. government. Ronald Eisler, Ph.D. Senior Research Biologist U.S. Geological Survey Patuxent Wildlife Research Center Laurel, Maryland


Acknowledgments This project was completed under the aegis of James A. Kushlan, director of the U.S. Geological Survey Patuxent Wildlife Research Center (PWRC), and Richard L. Jachowski, chief scientist at PWRC. I am grateful for their support and encouragement. Other research managers and scientists at PWRC who contributed significantly to the project over the 17 years that it was operative — and to whom I am indebted — include Peter H. Albers, Andre A. Belisle, W. Nelson Beyer, Lawrence J. Blus, Donald R. Clark, Jr., Nancy C. Coon, Harry N. Coulombe, Thomas W. Custer, Lawrence R. DeWeese, W. James Fleming, John B. French, Jr., Martha L. Gay, Christopher E. Grue, Russell J. Hall, Gary H. Heinz, Charles J. Henny, Paula F. P. Henry, Elwood F. Hill, David J. Hoffman, Kirk A. King, Alex J. Krynitsky, Jerry R. Longcore, T. Peter Lowe, Mark J. Melancon, John F. Moore, Harold J. O’Conner, Harry M. Ohlendorf, Oliver H. Pattee, H. Randolph Perry, Jr., Matthew C. Perry, Barnett Rattner, Gregory J. Smith, Donald W. Sparling, Charles J. Stafford, Kenneth L. Stromborg, Douglas M. Swineford, Michael W. Tome, David L. Trauger, Donald H. White, and Stanley N. Wiemeyer. Computer and graphics assistance was kindly provided by PWRC specialists Kinard Boone, Henry C. Bourne, Lois M. Loges, and Robert E. Munro. Library services — a major component of this project — were expertly provided by Lynda J. Garrett, chief librarian at PWRC, and a number of assistant librarians, most notably P. Wanda Manning. I also acknowledge the technical assistance provided in early phases of this project by numerous scientists and editors from other U.S. Department of the Interior installations, especially researchers from the U.S. Geological Survey Columbia Environmental Research Center, and by scientists from the U.S. Environmental Protection Agency, the U.S. Department of Agriculture, various universities, the International Atomic Energy Agency, the Canadian Wildlife Service, and the U.S. Department of Energy.


CHAPTER

1

Cadmium 1.1

INTRODUCTION

There is no evidence that cadmium (Cd) is biologically essential or beneficial; on the contrary, it has been implicated as the cause of numerous human deaths and various deleterious effects in fish and wildlife. In sufficient concentration, it is toxic to all forms of life, including microorganisms, higher plants, animals, and man. It is a relatively rare metal, usually present in small amounts in zinc ores, and is commercially obtained as an industrial by-product of the production of zinc, copper, and lead. Major uses of cadmium are in electroplating, in pigment production, and in the manufacture of plastic stabilizers and batteries. Anthropogenic sources of cadmium include smelter fumes and dusts, the products of incineration of cadmium-bearing materials and fossil fuels, fertilizers, and municipal wastewater and sludge discharges; concentrations are most likely highest in the localized regions of smelters or in urban industrialized areas (Hammons et al. 1978; U.S. Environmental Protection Agency [USEPA] 1980; Nriagu 1980, 1981; Hutton 1983b; Eisler 1985; Scheuhammer 1987; U.S. Public Health Service [USPHS] 1993; Cooke and Johnson 1996). Industrial consumption of cadmium in the United States, estimated at 6000 metric tons in 1968, is increasing; projected use in the year 2000 is about 14,000 tons, primarily for electroplating of motor parts and in the manufacture of batteries. The cadmium load in soils and terrestrial biota in other industrialized countries also appears to be increasing and is of great concern in Scandinavia (Tjell et al. 1983), Germany (Markard 1983), and the United Kingdom (Hutton 1983a).

1.2

ENVIRONMENTAL CHEMISTRY

Cadmium, as cadmium oxide, is obtained mainly as a by-product during the processing of zincbearing ores and also from the refining of lead and copper from sulfide ores (USPHS 1993). In 1989, the United States produced 1.4 million kg of cadmium (usually 0.6 to 1.8 million kg) and imported an additional 2.7 million kg (usually 1.8 to 3.2 million kg). Cadmium is used mainly for the production of nickel–cadmium batteries (35%), in metal plating (30%), and for the manufacture of pigments (15%), plastics and synthetics (10%), and alloys and miscellaneous uses (10%) (USPHS 1993). Cadmium is a silver-white, blue-tinged, lustrous metal that melts at 321˚C and boils at 767˚C. This divalent element has an atomic weight of 112.4, an atomic number of 48, and a density of 8.642 g/cm3. It is insoluble in water, although its chloride and sulfate salts are freely soluble (Windholz et al. 1976; USPHS 1993). The availability of cadmium to living organisms from their immediate physical and chemical environs depends on numerous factors, including adsorption and desorption rates of cadmium from terrigenous materials, pH, Eh, chemical speciation, and many


other modifiers. The few selected examples that follow demonstrate the complex behavior of cadmium in aquatic systems. Microbial extracellular polymeric substances (EPS) — ubiquitous features in aquatic environments — actively participate in binding dissolved overlying and pore-water metals in sediments. Organic sediment coatings in the form of bacterial EPS equivalent to about 0.5% organic matter can adsorb cadmium under estuarine conditions (Schlekat et al. 1998). EPS aggregates rapidly sorbed up to 90% of cadmium from solution. Changes in pH affected cadmium sorption, with the proportion of freed Cd to sorbed Cd changing from 90% at pH 5 to 5% at pH 9; desorption was enhanced with increasing salinity (Schlekat et al. 1998). Adsorption and desorption processes are likely to be major factors in controlling the concentration of cadmium in natural waters and tend to counteract changes in the concentration of cadmium ions in solution (Gardiner 1974). Adsorption and desorption rates of cadmium are rapid on mud solids and particles of clay, silica, humic material, and other naturally occurring solids. Concentration factors for river muds varied between 5000 and 500,000 and depended mainly on the type of solid, the particle size, the concentration of cadmium present, the duration of contact, and the concentration of complexing ligands; humic material appeared to be the main component of river mud responsible for adsorption (Gardiner 1974). Changes in physicochemical conditions, especially pH and redox potential, that occur during dredging and disposal of cadmium-polluted sediments may increase chemical mobility and, hence, bioavailability of sediment-bound cadmium (Khalid et al. 1981). For example, cadmium in Mississippi River sediments spiked with radiocadmium was transformed from potentially available organic forms to more mobile and readily available dissolved and exchangeable forms (i.e., increased bioavailability) under regimens of comparatively acidic pH and high oxidation (Khalid et al. 1981). The role of dissolved oxygen and aquatic plants on cadmium cycling was studied in Palestine Lake, a 92-ha eutrophic lake in Kosciusko County, Indiana, a longterm recipient of cadmium and other waste metals from an electroplating plant. The maximum recorded concentration of dissolved cadmium in the water column was 17.3 µg/L; for suspended particulates, it was 30.3 µg/L (Shephard et al. 1980). During anaerobic conditions in the lake’s hypolimnion, a marked decrease in the dissolved fraction and a corresponding increase in the suspended fraction were noted. The dominant form of cadmium was free, readily bioavailable, cadmium ion, Cd2+; however, organic complexes of cadmium, which are comparatively nonbioavailable, made up a significant portion of the total dissolved cadmium. Cadmium levels in sediments of Palestine Lake ranged from 1.5 mg/L in an uncontaminated area of the lake to 805 mg/L near the outlet of a metal-bearing ditch that entered the lake (McIntosh et al. 1978). The dominant form of cadmium in sediments was as a carbonate. Levels of cadmium in water varied over time and between sites, but usually ranged from 0.5 to 2.5 µg/L. It is possible that significant amounts of cadmium are transferred from the sediments into rooted aquatic macrophytes and later released into the water after macrophyte death (natural or herbicide-induced), particularly in heavily contaminated systems. In Palestine Lake, cadmium levels in pondweed (Potomogeton crispus), a rooted aquatic macrophyte, were about 90 mg/kg dry weight; a maximum burden of 1.5 kg was retained by the population of P. crispus in the lake (McIntosh et al. 1978). Release of the total amount could raise water concentrations by a maximum of 1 µg/L. This amount was considered negligible in terms of the overall lake cadmium budgets; however, it might have limited local effects.

1.3

CONCENTRATIONS IN FIELD COLLECTIONS

Small amounts of cadmium enter the environment from the natural weathering of minerals, but most is released as a result of human activities such as mining, smelting, fuel combustion, disposal of metal-containing products, and application of phosphate fertilizers or sewage sludges (USPHS 1993). In 1988, an estimated 306,000 kg of cadmium entered the domestic environment as a result


Table 1.1

Cadmium Burdens and Residence Times in the Principal Global Reservoirs

Reservoir Atmosphere Oceans Dissolved Suspended particulates (total) Particulate organic matter Freshwaters Dissolved Sediments Groundwater Glaciers Sediment pore waters Swamps and marshes, biomass Biosphere Marine plants Marine animals Land plants Land animals Freshwater biota Human biomass Terrestrial litter Lithosphere (down to 45 km)

Concentration 0.00003 µg/m3

Total Cd in Reservoir, in Metric Tons 1500

Residence Time 7 days

0.06 µg/kg 1.0 mg/kg 4.5 mg/kg

84,000,000 1,400,000 320,000

21,000 years — 1.3 years

0.05 µg/kg 0.16 mg/kg 0.1 µg/kg 0.005 µg/kg 0.2 µg/kg 0.6 mg/kg

16,000 100,000 400 82,000 64,000,000 3600

— 3.6 years — — — —

2.0 mg/kg 4.0 mg/kg 0.3 mg/kg 0.3 mg/kg 3.5 mg/kg 50 mg/person 0.6 mg/kg 0.5 mg/kg

400 12,000 720,000 6000 7000 200 1,300,000 2.8 × 1013

18 days — 20 days — 3.5 years 1–40 years 42 years 1 billion years

Adapted from Nriagu, J.O. (ed.) 1980. Cadmium in the Environment. Part I: Ecological Cycling. John Wiley, NY. 682 pp.

of human activities, mostly from industrial releases that were subsequently applied to soils (USPHS 1993). Where cadmium is comparatively bioavailable, these values are very near those that have been shown to produce harmful effects in sensitive biological species, as will be discussed later. Loadings of cadmium in uncontaminated, nonbiological compartments extended over several orders of magnitude, with most of the cadmium in lithosphere and ocean reservoirs (Table 1.1; Korte 1983). Concentrations (µg/L or µg/kg) of cadmium reported in uncontaminated compartments ranged from 0.05 to 0.2 in freshwater, up to 0.05 in coastal seawater, from 0.01 to 0.1 in openocean seawater, 0.1 to 14 in stormwater runoff, up to 5000 in riverine and lake sediments, 30 to 1000 in marine sediments, 10 to 1000 in soils of nonvolcanic origin, as much as 142 in human dietary items, up to 4500 in soils of volcanic origin, 1 to 600 in igneous rock, up to 100,000 in phosphatic rock, and 0.001 to 0.005 µg/m 3 in air (Korte 1983; USPHS 1993). Higher values are reported in abiotic materials and living organisms from cadmium-contaminated environments, being highest near cadmium-emitting industries such as smelters, municipal incinerators, and fossil fuel combustion facilities (Eisler 1985; USPHS 1993). Cadmium, unlike synthetic compounds, is a naturally occurring element, and its presence has been detected in more than 1000 species of aquatic and terrestrial flora and fauna. Concentrations of cadmium in selected species of biota are shown in Table 1.2; more extensive documentation was presented by Hammons et al. (1978), NRCC (1979), Jenkins (1980), and Eisler (1981). At least six trends are evident from Table 1.2. First, marine biota generally contained significantly higher cadmium residues than their freshwater or terrestrial counterparts, probably because total cadmium levels are higher in seawater. Second, cadmium tends to concentrate in the viscera of vertebrates, especially the liver and kidneys. Third, concentrations of cadmium are higher in older organisms than in younger stages; this relationship is especially pronounced in carnivores and marine vertebrates (Eisler 1984). Fourth, higher concentrations reported for individuals of a single species collected at several locations are almost always associated with proximity to industrial and urbanized


areas or to point source discharges of cadmium-containing wastes. Fifth, background levels of cadmium in crops and other plants are usually 90% for naphthalene in 24 h, and 20% (muscle), to 90% (gill) for benzo[a]pyrene in a similar period (Lee et al. 1972). Phenanthrene is metabolized by many species of aquatic organisms, including fish. A marine flounder Platichthys flesus, given a single oral dose of 0.7 mg phenanthrene/kg body weight, contained elevated phenanthrene concentrations in lipids, melanin-rich tissues (such as skin), and the eye lens; most was eliminated within 2 weeks (Solbakken et al. 1984). Studies on uptake rates of radiolabeled phenanthrene by marine polychaetes (Aberenicola pacifica) from PAH-contaminated sediments of high or low organic content showed that rates increased when worms were acclimatized for 48 h to sediments of low organic content (Penry and Weston 1998). Different rates of accumulation and depuration of benzo[a]pyrene and naphthalene in bluegill (Lepomis macrochirus) and Daphnia magna have been documented by McCarthy and Jimenez (1985) and McCarthy et al. (1985). Benzo[a]pyrene accumulations in bluegill, for example, were 10 times greater than naphthalene, but benzo[a]pyrene is extensively metabolized, whereas naphthalene is not. Consequently, postexposure accumulations of naphthalene greatly exceeded that of the parent benzo[a]pyrene. Because the more hydrophobic PAHs, such as benzo[a]pyrene, show a high affinity for binding to dissolved humic materials and have comparatively rapid biotransformation rates, these interactions may lessen or negate bioaccumulation and food chain transfer of hydrophobic PAHs (McCarthy and Jimenez 1985; McCarthy et al. 1985). Time to depurate or biotransform 50% of accumulated PAHs (Tb 1/2) varied widely. Tb 1/2 values for Daphnia pulex and all PAH compounds studied ranged between 0.4 and 0.5 h (Southworth et al. 1978). In rainbow trout given an intraarterial injection of 10 mg/kg BW of 2-methylnaphthalene, fluorene, or pyrene, the Tb 1/2 values ranged between 9.6 and 12.8 h; when route of exposure was intragastric and doses were 50 mg/kg BW, there was negligible uptake (Kennedy and Law 1990). For marine copepods and naphthalene, a Tb 1/2 of about 36 h was recorded (Neff 1982a).


For most marine bivalve molluscs, Tb 1/2 values ranged from 2 to 16 days. Some species, such as the hardshell clam (Mercenaria mercenaria), showed little or no depuration, while others, such as oysters, eliminated up to 90% of accumulated PAHs in 2 weeks — although the remaining 10% was released slowly and traces may remain indefinitely (Jackim and Lake 1978). Percent loss of various PAHs in American oysters (Crassostrea virginica), 7 days postexposure, ranged from no loss for benzo[a]pyrene to 98% for methylnaphthalene; intermediate were benz[a]anthracene (32%), fluoranthene (66%), anthracene (79%), dimethylnaphthalene (90%), and naphthalene (97%) (Neff 1982a). Teleosts and arthropods usually had low Tb 1/2 values. In bluegill, 89% loss of benzo[a]pyrene was recorded 4 h postexposure. For midge larvae, it was 72% in 8 h; and for daphnids, it was 21% in 18 h (Leversee et al. 1981). The role of sediments in PAH uptake kinetics should not be discounted (Hontela et al. 1992; van der Weiden et al. 1993; Hellou et al. 1994b). Yellow perch (Perca flavescens) and northern pike (Esox lucius) exposed to sediments contaminated with up to 3.7 mg PAHs/kg had pituitary damage and were unable to increase serum cortisol in response to the acute stress of capture when compared to conspecifics from uncontaminated sites, suggesting that life-long exposure to PAHs will destabilize the cortisol-producing endocrine system (Hontela et al. 1992). Male winter flounders (Pleuronectes americanus) exposed to PAH-contaminated sediments at concentrations as low as 1 mg/kg for 4 months had altered mixed-function oxygenase levels and fat content of liver (Payne et al. 1988). Sediment-associated anthracene contributed about 77% of the steady-state body burden of this compound in the amphipod Hyalella azteca (Landrum and Scavia 1983). For benzo[a]pyrene and the amphipod Pontoporeia hoyi, the sediment source (including interstitial water) accounted for 53% in amphipods collected at 60 m, but only 9% at 23 to 45 m (Landrum et al. 1984). Benthos from the Great Lakes, such as oligochaete worms (Limnodrilus sp., Stylodrilus sp.) and amphipods (Pontoporeia hoyi), obtain a substantial fraction of their PAH body content from the water when sediment PAH concentrations are low. However, when sediment PAH concentrations are elevated, benthos obtain a majority of their PAHs from that source through their ability to mobilize PAHs from the sediment/pore water matrix; the high concentrations of phenanthrene, fluorene, benzo[a]pyrene, and other PAHs measured in these organisms could provide a significant source of PAHs to predator fish (Eadie et al. 1983). Great Lakes benthos appear to contain as much PAHs as the fine-grain fraction of the sediment that serves as their food, although overlying water or pore water appears to contribute a larger proportion of PAHs to the organism’s body burden than do sediments (Eadie et al. 1984). Marine mussels (Mytilus edulis) and polychaete annelid worms (Nereis virens) exposed for 28 days to sediments heavily contaminated with various PAH compounds accumulated significant concentrations (up to 1000 times control levels) during the first 14 days of exposure, and little thereafter; during a 5-week postexposure period, depuration was rapid, with the more water-soluble PAHs excreted most rapidly; PAH levels usually remained above control values to the end of the postexposure period (Lake et al. 1985). English sole (Pleuronectes vetulus), during exposure for 11 to 51 days to PAH-contaminated sediments, showed significant accumulations of naphthalenes in liver (up to 3.1 mg/kg dry weight) after 11 days, with concentrations declining markedly thereafter; uptake of phenanthrene, chrysene, and benzo[a]pyrene was negligible during the first 7 days (Neff 1982a). Selected studies on effects of benzo[a]pyrene to aquatic vertebrates are summarized in Table 25.7 and demonstrate dose-dependent increases in death, enzyme disruption, liver and intestinal histopathology, carcinogenic and mutagenic effects, and behavioral deficits. In general, effects of benzo[a]pyrene were exacerbated by poor health, and by interactions with radiation or cadmium; effects were modified by route of exposure, diet, and organism age (Table 25.7). Exposure of fishes to BaP at early life stages has potential life-long significance. Embryonic sublethal doses of BaP are capable of inducing sublethal changes in behavior weeks or even months following hatching of salmon and trout (Ostrander et al. 1988, 1990). Fluorene effects in freshwater pond ecosystems have been evaluated (Boyle et al. 1984, 1985; Finger et al. 1985). In ponds exposed to initial fluorene concentrations of 0.12 to 2.0 mg/L, Tb 1/2


Table 25.7

Benzo[a]pyrene Effects on Selected Aquatic Vertebrates Effect

Referencea

BaP alone produced glycogen depletion in liver; cadmium alone caused hepatic perivascular fibrosis. Mixture produced complete disorganization of the hepatic parenchyma, including nuclear degeneration; higher increase in EROD activity (by 19-fold), BaP hydroxylase activity (by 71-fold), and cytochrome P-450 microsomal content by 2-fold

1

Organism, Dose, and Other Variables EUROPEAN EEL, Anguilla anguilla Seawater-adapted eels were exposed for 24 days to 0 or 5 µg cadmium/L, then given a single intraperitoneal (ip) injection of 20 mg BaP/kg BW and killed 24 h later

WHITE SUCKER, Catostomus commersonii Suckers with diseased bile ducts were given a single oral dose of BaP

Bile duct disease does not appear to restrict the metabolism and excretion of BaP. However, BaP metabolite levels were 57% higher in livers of affected suckers than controls, and may be due to bile retention within the diseased bile ducts

2

After 72 h, bile contained 25% of the total radioactivity; major BaP metabolites were glucuronides (54%), sulfates (12%), and unmetabolized BaP(14%). The potentially genotoxic metabolite BaP-7,8-dihydrodiol and its glucuronide represented 0.7 and 2.0%, respectively, of the bile radioactivity BaP and chrysene were most efficient in elevating EROD activities, cytochrome P-4501A protein levels, and total cytochrome P-450 content 1–14 days postinjection. Less pronounced increases were caused by fluoranthene and pyrene; B[ghi]PER did not affect these parameters

3

COMMON CARP, Cyprinus carpio Single ip injection of radiolabeled BaP

Single ip injection of 99.5 µmol/kg BW of BaP, chrysene, fluoranthene, pyrene, or benzo[ghi]perylene

4

EUROPEAN SEA BASS, Dicentrarchus labrax Single ip injection of radiolabeled BaP

Immatures given single ip injection of 20 mg/kg BW, or by force feeding 20 mg/kg BW. Liver and intestine examined for pathology 17 days after treatment Force feeding radiolabeled BaP to immatures; single administration Fed diets containing fish oil vs. no fish oil diet for 5 months, then given single ip injection of BaP. Hepatic transformation enzymes measured 14 days postinjection

Half-time persistence (Tb 1/2) ranged from 2.2–2.4 days for liver, gills, gonads, and whole body; 2.9–5.1 days for muscle, spleen, gall bladder, and intestine; 6.5 days for kidney; and 12.4 days for fat Injection route of intoxication produced severe liver histopathology when compared to dietary route; intestinal histopathology observed with dietary route but not with injection route Gallbladder had 85% of radioactivity. Tb 1/2 values were 0.8 days for liver, 3.3 days for intestine, 3.5 days for gallbladder, and 8.2 days for liver Diets rich in fish oils produced marked increases in activities of P-450 cytochromes, 7-EROD, and other enzyme activities; bass fed diets devoid of fish oil had lower microsomal enzyme activities

5

6

7

8

GIZZARD SHAD, Dorosoma cepedianum Single ip injection equivalent to 0, 0.1, 1, 10, or 50 mg BaP/kg BW and killed 72 h later Held for 10–20 days in water containing 0, 0.14, 0.24, 0.44, or 0.76 µg BaP/L


Maximum EROD activity was reached in the 10 and 50 mg/kg groups; no effect on EROD activity at 1 mg/kg BW and lower After 10 days, no effect on EROD enzyme activity at 0.24 µg/L and lower; EROD activity significantly elevated at 0.44 µg/L and higher. Concentration of parent BaP in whole fish after 20 days ranged from 20 µg/kg FW in the 0.14 µg/L group to 26.5 µg/kg FW in the 0.76 µg/L group

9

9

Table 25.7 (continued)

Benzo[a]pyrene Effects on Selected Aquatic Vertebrates Effect

Referencea

Dose-dependent increase in liver BaP monooxygenase activity; adverse effect on liver DNA at high dose

10

Organism, Dose, and Other Variables MOSQUITOFISH, Gambusia affinis Exposed for 48 h in water containing 1–100 µg BaP/L

BROWN BULLHEAD, Ictalurus nebulosus Subadults given single ip injection of radiolabeled BaP and killed 72 h later

Adults given single ip injection of 0, 5, 25, or 125 mg BaP/kg BW and observed for 18 months

Biliary radioactivity accounted for 16% of the radioactivity; the potentially genotoxic metabolite BaP-7,8-dihydrodiol accounted for 3% of the bile radioactivity. Other accumulation sites were muscle, liver, and kidney At 18 months, dose-dependent changes in liver pigmentation; no evidence of long-term DNA alterations; no neoplasms detected

11

Hypomethylation, as judged by loss of 5-methyl deoxycytidine content from DNA, occurred shortly after BaP exposure and continued upon termination of exposure conditions. Onset and persistence of hypomethylation were correlated with other types of DNA-damaging events, including strand breaks and DNA adduct formation BaP was not detectable in muscle and gut, indicating that BaP on sediments may be mobilized and made available to benthic invertebrates and fish but that the process was highly inefficient with minor biomagnification; a similar case is made for fluoranthene

13

12

BLUEGILL, Lepomis macrochirus Held for 40 days in 1 µg BaP/L, then for 45 days in uncontaminated media

Chironomids (Chironomus riparius) held for 72 h on sediments with 47–4040 µg BaP/kg contained a maximum of 6030 µg BaP/kg DW; chironomids were fed to bluegills for 96 h

14

SUNFISH, Lepomis spp. hybrids Single ip injection of 8 mg BaP/kg BW and killed 4–5 days later

Hepatotoxicity. Decreased glutathione S-transferase activity in liver

15

RAINBOW TROUT, Oncorhynchus mykiss Juveniles given single intramuscular (im) injection of radiolabeled BaP Two days prior to hatch fertilized eggs were given a single transchorionic injection of 0.09–18.0 µg BaP/egg, equivalent to 1.2–197.8 mg BaP/kg FW egg. Mortality and neoplasms documented over the next 12 months

Newly fertilized embryos were incubated in water containing 0, 0.08, 0.2, 0.4, 1.5, 2.4, or 3.0 µg BaP/L. Alevins examined

Subadults exposed to 0.4 or 1.2 µg BaP/L for 30 days Subadults exposed to 0.5 µg BaP/L for 15 days then a pulse of 60 µg/L which naturally declined to 2.0 µg/L for another 15 days


After 48 h, liver DNA and RNA contained 1–2.4% of administered dose; a similar case is made for 7,12dimethylbenz[a]anthracene Control mortality ranged between 6.5 and 10%; mortality in BaP-treated fish ranged between 40 and 78% in a dose-dependent manner. Liver neoplasm frequency in controls at age 1 year was 1.8%; for the 4.5–9.0-µg groups (49.5–91.0 mg/kg), these values ranged between 5 and 6%; for the 18-µg group (197.8 mg/kg FW), the liver neoplasm frequency was 16% at age 12 months after a single-dose exposure as embryos Histologic and skeletal abnormalities in all BaPexposed groups; depressed mitotic rates in retina and brain in groups exposed to 0.08 µg/L and higher; muscle necrosis in all groups exposed to 0.2 µg/L and higher; microphthalmia was observed in 17% of BaP-treated trout No DNA adducts detected in liver

16

19

Liver DNA adducts detected

19

17

18


Benzo[a]pyrene Effects on Selected Aquatic Vertebrates

Organism, Dose, and Other Variables Subadults given single ip injection of 20 mg/kg BW and killed after 7 days Subadults given two ip injections of 25 mg/kg BW separated by 7 days and killed 14 days after the second injection Subadults fed diets for 6 days containing either 58 or 288 µg BaP daily, equivalent to 43 mg/kg BW or 216 mg/kg BW Embryos exposed for 24 h to 25 mg BaP/L, one week prior to hatch

Effect

Referencea

No DNA adducts found in liver

19

DNA adducts found in liver

19

Liver DNA adducts detected in both cases

19

No effect on hatch, but alevins showed decreased tendency to orient upstream

20

COHO SALMON, Oncorhynchus kisutch Fertilized eggs exposed to 7, 10, or 25 mg BaP/L for 24 h at 1 week prior to hatch or 24 h after fertilization

BaP was easily accumulated and rapidly lost. Altered hatching rate and reduced survival at higher concentrations. Significant behavioral alterations in the months following hatching, including difficulty in upstream orientation, swimming, and foraging ability; fry were the most sensitive stage

21, 22

Dose-dependent increase in metabolite production. Rapid metabolism of BaP independent of temperature Accumulation followed a dose-concentration gradient; uptake rate was higher for cells acclimatized to 18°C than those acclimatized to 28°C at all incubation temperatures. When cells were exposed to BaP at the respective acclimatization temperatures, uptake rates were similar

23

Dose-dependent reduction in pronephros cell counts

25

Liver neoplasms at intermediate and high concentration exposures

26

Visible light alone, UV radiation alone, or BaP alone had no toxic effects. Genotoxic effects were observed at 500 µg BaP/L plus visible light, or 25 µg BaP/L plus UV radiation

27

GULF TOADFISH, Opsanus tau Isolated hepatocytes acclimatized to 18 or 28°C, then exposed to BaP at 18, 23, or 28°C for about 2 h Isolated gill cells acclimatized to either 18 or 28°C for 3 weeks and exposed for 8 h to 1–100 mg BaP/L at incubation temperatures of 18, 23, and 28°C

24

NILE TILAPIA, Oreochromis niloticus Single ip injection, various doses MEDAKA, Oryzias latipes Exposed to approximately 5, 40, or 200 µg BaP/L for 6 h, once weekly, for up to 4 weeks; 6–10-day-old fish NEWT, Pleurodeles waltl Embryos and larvae exposed to visible light alone, ultraviolet radiation alone, BaP alone at 12.5–500 µg/L, or combinations

WINTER FLOUNDER, Pleuronectes americanus Single oral dose of radiolabeled BaP and killed 8–96 h after dosing

After 8 h, 8.1% of the original dose administered was in intestine, 3.9% in blood, 1.6% in liver, 0.6% in bile, and 1.0 mg/kg of acenaphthalene (2.4), phenanthrene (5.7), fluoranthene (1.9), and pyrene (1.1), and lower concentrations of heavier molecular weight PAHs. Bullheads also exhibited a high (33%) liver tumor frequency, which seemed to correspond to their PAH body burdens. Investigators concluded that the elevated frequency of liver neoplasia in Black River bullheads was chemically induced and was the result of exposure to PAHs (Baumann et al. 1982; Baumann and Harshbarger 1985). Neoplasms in several species of fishes have been produced experimentally with 3-methylcholanthrene, acetylaminofluorene, benzo[a]pyrene, and 7,12-dimethylbenz[a]anthracene, with tumors evident 3 to 12 months postexposure (Couch and Harshbarger 1985; Hendricks et al. 1985; Hawkins et al. 1989). Under laboratory conditions, liver neoplasms were induced in two species of minnows (Poeciliopsis spp.) by repeated short-term exposures (6 h once a week, for 5 weeks) to an aqueous suspension of 5 mg/L of 7,12-dimethylbenz[a]anthracene. About 44% of the fish surviving this treatment developed hepatocellular neoplasms 6 to 9 months postexposure (Schultz and Schultz 1982). Guppies developed hepatic and extrahepatic neoplasms within 6 months following brief (6-h exposure, once weekly for 4 weeks) waterborne exposures to very low (20 to 35 µg/L) concentrations of 7,12-dimethylbenz[a]anthracene (Hawkins et al. 1989). Eastern mudminnows (Umbra pygmaea) kept in water containing up to 700 µg PAHs/L for 11 days showed increased frequencies of chromosomal aberrations in gills: 30% vs. 8% in controls (Prein et al. 1978). High dietary benzo[a]pyrene levels of 500 mg/kg produced significant elevations in hepatic mixed-function oxidase levels in rainbow trout after 9 weeks (Hendricks et al. 1985). Rainbow trout fed diets containing 1006 mg benzo[a]pyrene/kg for 12 months developed liver tumors (Couch et al. 1983). About 25% of rainbow trout kept on diets containing 1000 mg benzo[a]pyrene/kg for 18 months had histologically confirmed liver neoplasms as compared to 15% after 12 months, with no evidence of neoplasia in controls (Hendricks et al. 1985). Young English sole may activate and degrade carcinogenic PAHs, such as benzo[a]pyrene, to a greater extent than adults, but additional research is needed to determine if younger fish are at greater risk than older sole to PAH-induced toxicity (Varanasi et al. 1984). In English sole, a high significant positive correlation between PAH metabolites (1- and 3-hydroxy benzo[a]pyrene, hydroxy and dehydrodiol metabolites of pyrene and fluoranthene) in bile, and idiopathic liver lesions, prevalence of neoplasms, megalocytic hepatosis, and total number of hepatic lesions (Krahn et al. 1986) suggests that selected PAH metabolites and key organs or tissues may be the most effective monitors of PAH contamination in aquatic organisms. In addition to those effects of PAHs emphasizing survival, uptake, depuration, and carcinogenesis previously listed, a wide variety of additional effects have been documented for aquatic organisms. These include: • Inhibited reproduction of daphnids and delayed emergence of larval midges by fluorene (Finger et al. 1985) • Reproductive impairment in fish by anthracene (Hall and Oris 1991) and elevated concentrations of total PAHs in sediments (Johnson et al. 1998) • Generalized disruption of cell membrane function in fishes by anthracene (McCloskey and Oris 1993) • Decreased respiration and heart rate in mussels (Mytilus californianus) by benzo[a]pyrene (Sabourin and Tullis 1981) • Increased weight of liver, kidney, gall bladder, and spleen of sea catfish (Arius felis) by 3-methylcholanthrene, which was dose-related (Melius and Elam 1983) • Photosynthetic inhibition of algae and macrophytes by anthracene, naphthalene, phenanthrene, pyrene (Neff 1985), and fluorene (Finger et al. 1985) • Immobilization of the protozoan Paramecium caudatum by anthracene, with an EC50 (60 min) of 0.1 µg/L) (USEPA 1980) • Perylene accumulation by algae (Stegeman 1981)


• Accumulation without activation of benzo[a]pyrene and benzo[a]anthracene by a marine protozoan (Parauronema acutum), and biotransformation of various fluorenes by P. acutum to mutagenic metabolites (Lindmark 1981) • Interference by toluene and anthracene with benzo[a]pyrene uptake by freshwater amphipods (Landrum 1983) • Abnormal blood chemistry in oysters (Crassostrea virginica) exposed for 1 year to 5 µg 3-methylcholanthrene/L (Couch et al. 1983) • Enlarged livers in brown bullheads from a PAH-contaminated river (Fabacher and Baumann 1985).

25.4.5 Amphibians and Reptiles In vitro studies with abdominal skin of Rana pipiens demonstrated that naphthalene inhibited sodium transport after exposure to 4.4 mg/L for 30 min (Blankemeyer and Hefler 1990). Some data were available on biological effects of benzo[a]pyrene, 3-methylcholanthrene, and perylene to reptiles and amphibians (Balls 1964; Stegeman 1981; Anderson et al. 1982; Schwen and Mannering 1982a, 1982b; Couch et al. 1983). Implantation of 1.5 mg benzo[a]pyrene crystals into the abdominal cavity of adult South African clawed frogs (Xenopus laevis) produced lymphosarcomas in 11 of the 13 frogs (85%) after 86 to 288 days (Balls 1964). Immature frogs were more resistant, with only 45% bearing lymphoid tumors of liver, kidney, spleen, or abdominal muscle 272 to 310 days after implantation of 1.5 mg of benzo[a]pyrene crystals in the dorsal lymph sac or abdominal cavity. Implantation of 3-methylcholanthrene crystals into X. laevis provokes development of lymphoid tumors similar to those occurring naturally in this species. Moreover, these tumors are readily transplantable into other Xenopus or into the urodele species Triturus cristatus (Balls 1964). Intraperitoneal injection of perylene into tiger salamanders can result in hepatic tumors (Couch et al. 1983). A critical point of interaction between PAHs and reptiles or amphibians involves the transformation of these compounds by cytochrome P-450-dependent monooxygenase systems (Stegeman 1981); in general, reaction rates in this group are considerably slower than those observed in hepatic microsomes from mammals (Schwen and Mannering 1982a). Mixed-function oxidation systems can be induced in liver and skin of tiger salamanders by perylene (Couch et al. 1983) and 3-methylcholanthrene (Anderson et al. 1982), and in liver of the leopard frog (Rana pipiens) and garter snake (Thamnophis sp.) by benzo[a]pyrene and 3-methylcholanthrene (Stegeman 1981; Schwen and Mannering 1982a, 1982b). A single dose of 40 mg/kg body weight of 3-methylcholanthrene was sufficient to induce mixed-function oxidase activity for several weeks in the leopard frog (Schwen and Mannering 1982b). Amphibians, including tiger salamanders, are quite resistant to PAH carcinogenesis when compared to mammals, according to Anderson et al. (1982). This conclusion was based on studies with Ambystoma hepatic microsomes and their inability to produce mutagenic metabolites of benzo[a]pyrene and perylene (as measured by bacterial Salmonella typhimurium strains used in the Ames test). However, rat liver preparations did produce mutagenic metabolites under these procedures (Anderson et al. 1982). 25.4.6 Birds In birds, PAHs were associated with impaired reproduction, growth retardation, morphological abnormalities, behavioral changes, and alterations in Vitamin A and thyroid hormone metabolism (Murk et al. 1996). In one study, fertilized eggs of the chicken (Gallus domesticus) and the common eider duck (Somateria mollissima) were injected with a mixture of 16 PAHs at 0.2 mg PAH/kg egg FW on day 4 of incubation (Naf et al. 1992). In chicken eggs, 94% of the administered dose was metabolized within the egg by day 18. In embryos of both chickens and eiders, PAH concentrations were highest in gallbladder, followed by liver, kidney, and adipose tissues. Eiders had 40% of the total PAH content in these organs vs. 16% for chickens. Chick embryos, eider embryos, and


juvenile eiders had similar PAH concentrations and PAH profiles. The largest PAH molecule studied, coronene, was not taken up from the yolk by the embryo as efficiently as other PAHs, but once taken up it was metabolized as readily as the other PAHs. AHH activities of chick and eider embryos were of similar magnitude (Naf et al. 1992). Studies with European starlings (Sturnus vulgaris) and 7,12-dimethylbenz[a]anthracene show that effects on the immune function and hepatic mixedfunction oxygenase activity were more pronounced when administered by subcutaneous injection when compared to an oral route of exposure, and more pronounced in nestlings than in adults. Serious adverse effects were noted in adults at total administered doses of 125 mg/kg BW via injection or orally, and 100 mg/kg BW in nestlings (Trust et al. 1994). Mallards (Anas platyrhynchos) fed diets that contained 4000 mg PAHs/kg (mostly as naphthalenes, naphthenes, and phenanthrene) for a period of 7 months had normal survival and no visible signs of toxicity during exposure. However, liver weight increased 25% and blood flow to liver increased 30%, when compared to controls (Patton and Dieter 1980). In another study with mallards, Hoffman and Gay (1981) measured embryotoxicity of various PAHs applied externally, in a comparatively innocuous synthetic petroleum mixture, to the surface of mallard eggs. The most embryotoxic PAH tested was 7,12-dimethylbenz[a]anthracene: approximately 0.002 µg/egg (equivalent to about 0.036 µg/kg fresh weight, based on an average weight of 55 g per egg) caused 26% mortality in 18 days, and, among survivors, produced significant reduction in embryonic growth and a significant increase in the percent of anomalies (e.g., incomplete skeletal ossification, defects in eye, brain, liver, feathers, and bill). At 0.01 µg 7,12-dimethylbenz[a]anthracene/egg, only 10% survived to day 18. Similar results were obtained with 0.015 µg (and higher) chrysene/egg. For benzo[a]pyrene, 0.002 µg/egg did not affect mallard survival, but did cause embryonic growth reduction and an increased incidence of abnormal survivors. At 0.01 µg benzo[a]pyrene/egg, 60% died in 18 days; at 0.05 µg/egg, 75% were dead within 3 days of treatment (Hoffman and Gay 1981). Embryos may contain microsomal enzymes that can metabolize PAHs to more highly toxic intermediates than can adults, and avian embryos may have a greater capacity to metabolize PAHs in this manner than do mammalian embryos and fetuses (as quoted in Hoffman and Gay 1981); this observation warrants additional research. Several investigators have suggested that the presence of PAHs in petroleum, including benzo[a]pyrene, chrysene, and 7,12-dimethylbenz[a]anthracene, significantly enhances the overall embryotoxicity in avian species, and that the relatively small percent of aromatic hydrocarbons contributed by PAHs in petroleum may confer much of the adverse biological effects reported after eggs have been exposed to microliter quantities of polluting oils (Hoffman and Gay 1981; Albers 1983). 25.4.7 Mammals Numerous PAH compounds are distinct in their ability to produce tumors in skin and in most epithelial tissues of practically all animal species tested. Malignancies were often induced by acute exposures to microgram quantities. In some cases, the latency period can be as short as 4 to 8 weeks, with the tumors resembling human carcinomas (USEPA 1980). Carcinogenic potential values of individual PAHs via dermal exposure, when compared to benzo[a]pyrene with an assigned value of 1.0, were 5.0 for dibenz[a,h]anthracene; 0.1 for benzo[a]anthracene, benzo[b,k]fluoranthene, and indeno[1,2,3-cd]perylene; 0.01 for benzo[g,h,i]perylene and chrysene; and 0.001 for fluoranthene, fluorene, phenanthrene, and pyrene (Hussain et al. 1998). Certain carcinogenic PAHs are capable of passage across skin, lungs, and intestine, and can enter the rat fetus, for example, following intragastric or intravenous administration to pregnant dams (USEPA 1980). In most cases, the process of carcinogenesis occurs over a period of many months in experimental animals, and many years in humans. The tissue affected is determined by the route of administration and species under investigation. Thus, 7,12-dimethylbenz[a]anthracene is a potent carcinogen for the mammary gland of young female rats after oral or intravenous administration; dietary benzo[a]pyrene leads


to leukemia, lung adenoma, and stomach tumors in mice, and both PAH compounds can induce hepatomas in skin of male mice when injected shortly after birth (Dipple 1985). Mammary tumors were observed in female rats 6 weeks after they were given the first of 4 weekly intragastric doses of 5 mg 7,12-dimethylbenz[a]anthracene/rat; exposure to magnetic fields of low flux density may promote the growth and development of mammary tumors (Mevissen et al. 1998). Acute and chronic exposure to various carcinogenic PAHs have resulted in destruction of hematopoietic and lymphoid tissues, ovotoxicity, antispermatogonic effects, adrenal necrosis, changes in the intestinal and respiratory epithelia, and other effects (Table 25.8) (USEPA 1980; Lee and Grant 1981). For the most part, however, tissue damage occurs at dose levels that would also be expected to induce carcinomas, and thus the threat of malignancy predominates in evaluating PAH toxicity. There is a scarcity of data available on the toxicological properties of PAHs that are not demonstrably carcinogenic to mammals (USEPA 1980; Lee and Grant 1981). Table 25.8

Some Effects of PAHs on Selected Laboratory Mammals

Effect (units), Organism, PAH Compound

Concentration

Referencea

LD50, ACUTE ORAL (mg/kg body weight) Rodents (Rattus spp., Mus spp.) Benzo[a]pyrene Phenanthrene Naphthalene Fluoranthene

50 700 1780 2000

1 1 1 1

CARCINOGENICITY, CHRONIC ORAL (mg/kg body weight) Rodents 7,12-Dimethylbenz[a]anthracene Benzo[a]pyrene Dibenz[a,h]anthracene Benz[a]anthracene Benzo[b]fluoranthene Benzo[k]fluoranthene Indeno[1,2,3-cd]pyrene Chrysene Anthracene

0.00004–0.00025 0.002 0.006 2.0 40.0 72.0 72.0 99.0 3300.0

2 1 1 1 1 1 1 1 1

CARCINOGENICITY, APPLIED EXTERNALLY AS TOPICAL (mg) Mice, Mus spp. Benzo[a]pyrene Dibenz[a,c]anthracene 7,12-Dimethylbenz[a]anthracene Dibenz[a,j]anthracene Anthracene Benzo[g,h,i]perylene Benz[a]anthracene

0.001 0.001 0.02 0.039 0.08 0.8 1.0

2 2 2 2 2 2 2

>0.0002 >0.00008

2 2

0.05 0.6 0.06 >0.6

2 2 2 2

CARCINOGENICITY, SUBCUTANEOUS (mg) Mice Dibenz[a,h]anthracene Adults Newborn Dibenzo[a,i]pyrene In sesame oil In peanut oil Benzo[a]pyrene Dibenzo[a,e]pyrene



Some Effects of PAHs on Selected Laboratory Mammals

Effect (units), Organism, PAH Compound Benzo[b]fluoranthene Benz[a]anthracene Dibenzo[a,h]pyrene

Concentration

Referencea

1.8 5.0 6.0

2 2 2

100.0 (no effect)

4

0.5–2.0 5.0 20.0

4 4 4

80.0 80.0 80.0

5 5 5

TESTICULAR DAMAGE (mg) Rat, Rattus spp. Benzo[a]pyrene, oral 7,12-Dimethylbenz[a]anthracene Intravenous Young rats Older rats Oral OOCYTE AND FOLLICLE DESTRUCTION, SINGLE INTRAPERITONEAL INJECTION (mg/kg body weight) Mice Benzo[a]pyrene 3-Methylcholanthrene 7,12-Dimethylbenz[a]anthracene

ALTERED BLOOD SERUM CHEMISTRY AND NEPHROTOXICITY, SINGLE INTRAPERITONEAL INJECTION (mg/kg body weight) Rat Phenanthrene Pyrene

150.0 150.0

6 6

FOOD CONSUMPTION, DAILY FOR 5 DAYS (mg/kg body weight) Deer mice, Peromyscus maniculatus 2-Methoxynaphthalene 30% reduction 2-Ethoxynaphthalene 3% reduction House mice, Mus musculus 2-Methoxynaphthalene 50% reduction 2-Ethyoxynaphthalene 50% reduction a

825

3

1213

3

825

3

1213

3

1, Sims and Overcash 1983; 2, Lo and Sandi 1978; 3, Schafer and Bowles 1985; 4, USEPA 1980; 5, Mattison 1980; 6, Yoshikawa et al. 1985.

Target organs for PAH toxic action are diverse, due partly to extensive distribution in the body and also to selective attack by these chemicals on proliferating cells (USEPA 1980). Damage to the hematopoietic and lymphoid system in experimental animals is a particularly common observation (USEPA 1980). In rats, the target organs for 7,12-dimethylbenz[a]anthracene are skin, small intestine, kidney, and mammary gland, whereas in fish the primary target organ is liver (Schultz and Schultz 1982). Application of carcinogenic PAHs to mouse skin leads to destruction of sebaceous glands and to hyperplasia, hyperkeratosis, and ulceration (USEPA 1980). Tumors are induced in mouse skin by the repeated application of small doses of PAHs, by a single application of a large dose, or by the single application of a subcarcinogenic dose (initiation) followed by repeated application of certain noncarcinogenic agents (promotion) (Dipple 1985). Newborn mice were highly susceptible to 3-methylcholanthrene, with many mice dying from acute or chronic wasting disease following treatment; some strains of mice eventually developed thymomas, but other strains showed no evidence despite serious damage to the thymus (USEPA 1980). In general, PAH


carcinogens transform cells through genetic injury involving metabolism of the parent compound to a reactive diol epoxide. This, in turn, can then form adducts with cellular molecules, such as DNA, RNA, and proteins, resulting in cell transformation (Dipple 1985; Ward et al. 1985). In the case of benzo[a]pyrene, one isomer of the 7,8-diol,9,10-epoxide is an exceptionally potent carcinogen to newborn mice and is believed to be the ultimate carcinogenic metabolite of this PAH (Slaga et al. 1978). One of the most toxicologically significant processes involved in the response to PAH absorption is the interaction with drug-metabolizing enzyme systems (Lee and Grant 1981). Increased production of mixed-function oxidase enzymes in various small mammals has been induced by halogenated naphthalenes (Campbell et al. 1983), 3-methylcholanthrene (Miranda and Chhabra 1980), and numerous other PAHs (USEPA 1980). PAH metabolites produced by microsomal enzymes in mammals can be arbitrarily divided into water-soluble groups, and organosoluble groups such as phenols, dihydrodiols, hydroxymethyl derivatives, quinones, and epoxides (USEPA 1980). In the case of benzo[a]pyrene, the diol epoxides are usually considered as the ultimate carcinogens. Other microsomal enzymes convert epoxide metabolites to easily excretable water soluble compounds, with excretion primarily through feces and the hepatobiliary system (USEPA 1980). Interspecies differences in sensitivity to PAH-induced carcinogenesis are due largely to differences in levels of fixed function oxidase activities, and these will directly affect rates at which active metabolites are converted to less-active products (Neff 1979). Investigators (Neff 1979; USEPA 1980; Dipple 1985) agree that: • Unsubstituted aromatic PAHs with fewer than 4 condensed rings have not shown tumorigenic activity • That many, but not all, four-, five-, and six-ring PAH compounds are carcinogenic • That only a few unsubstituted hydrocarbons with seven rings or more are tumorigenic or carcinogenic

Many PAH compounds containing four and five rings, and some containing six or more rings, provoke local tumors after repeated application to the dorsal skin of mice; the tumor incidence exhibited a significant dose-response relationship (Grimmer et al. 1985). Among unsubstituted PAHs containing a nonaromatic ring (e.g., cholanthrene and acenaphthanthracene), all active carcinogens retained an intact phenanthrene segment (USEPA 1980). The addition of alkyl substituents in certain positions in the ring system of a fully aromatic PAH will often confer carcinogenic activity or dramatically enhance existing carcinogenic potency. For example, monomethyl substitution of benz[a]anthracene can lead to strong carcinogenicity in mice, with potency depending on the position of substitution in the order 7 > 6 > 8 = 12 > 9; a further enhancement of carcinogenic activity is produced by appropriate dimethyl substitution, with 7,12-dimethylbenz[a]anthracene among the most potent PAH carcinogens known. Alkyl substitution of partially aromatic condensed ring systems may also add considerable carcinogenic activity as is the case with 3-methylcholanthrene. With alkyl substitutes longer than methyl, carcinogenicity levels decrease, possibly due to a decrease in transport through cell membranes (USEPA 1980). A good correlation exists between skin tumor-initiating activities of various benzo[a]pyrene metabolites and their mutagenic activity in mammalian cell mutagenesis systems (Slaga et al. 1978), although variations in chromosome number and structure may accompany tumors induced by various carcinogenic PAHs in rats, mice, and hamsters (Bayer 1978; USEPA 1980). Active PAH metabolites (e.g., dihydrodiols or diol epoxides) can produce sister chromatid exchanges in Chinese hamster ovary cells (Bayer 1978; USEPA 1980; Pal 1984). When exchanges were induced by the diol epoxide, a close relationship exists between the frequency of sister chromatid exchanges and the levels of deoxyribnucleoside–diol–epoxide adduct formation (Pal 1984). In general, noncarcinogenic PAHs were not mutagenic (USEPA 1980). Laboratory studies with mice have shown that many carcinogenic PAHs adversely affect the immune system, thus directly impacting an organism’s general health, although noncarcinogenic analogs had no immunosuppressive effect. Further, the more carcinogenic the PAH, the greater the immunosuppression (Ward et al. 1985). Destruction of


oocytes and follicles in mouse ovary is documented following intraperitoneal injection of benzo[a]pyrene, 3-methylcholanthrene, and 7,12-dimethylbenz[a]anthracene; the rate of destruction was proportional to the activity of the ovarian cytochrome P-450-dependent monooxygenase, as well as the carcinogenicity of the PAH (Mattison 1980). However, no information is presently available to indicate whether PAHs present a hazard to reproductive success. In those cases where teratogenic effects are clearly evident (e.g., 7,12-dimethylbenz[a]anthracene), the required doses were far in excess of realistic environmental exposures (Lee and Grant 1981). Numerous studies show that unsubstituted PAHs do not accumulate in mammalian adipose tissues despite their high lipid solubility, probably because they tend to be rapidly and extensively metabolized (USEPA 1980; Lee and Grant 1981). The biological half-life (Tb 1/2) of PAHs is limited, as judged by rodent studies. In the case of oral doses of benzo[a]pyrene and rat blood and liver, Tb 1/2 values of 5 to 10 min were recorded; the initial rapid elimination phase was followed by a slower disappearance phase lasting 6 h or more (USEPA 1980). The Tb 1/2 value of benzo[a]pyrene from blood of rats given 15 mg/kg BW by intravenous injection was 400 min (Moir et al. 1998). In that study, adipose and lung tissues had comparatively high concentrations of benzo[a]pyrene 32 h after dosing, and fecal excretion was the dominant route of BaP loss, being 4 to 10 times higher than urinary excretion (Moir et al. 1998). Tb 1/2 values from the site of subcutaneous injection in mice were 1.75 weeks for benzo[a]pyrene, 3.5 weeks for 3-methylcholanthrene, and 12 weeks for dibenz[a,h]anthracene; the relative carcinogenicity of each compound was directly proportional to the time of retention at the injection site (Pucknat 1981). Many chemicals are known to modify the action of carcinogenic PAHs in experimental animals, including other PAHs that are weakly carcinogenic or noncarcinogenic. The effects of these modifiers on PAH metabolism appear to fall into three major categories (DiGiovanni and Slaga 1981b): • Those that alter the metabolism of the carcinogen, causing decreased activation or increased detoxification • Those that scavenge active molecular species of carcinogens to prevent their reaching critical target sites in the cell • Those that exhibit competitive antagonism

For example, pyrene given to rats by intravenous injection or orally at 20 mg/kg BW alone or in combination with other PAHs produce the following observations: bioavailability of pyrene is increased in combination with fluoranthene or benz[a]anthracene; pyrene enhances the carcinogenic effect of benzo[a]pyrene, and carcinogenicity of benzo[a]pyrene can be inhibited by pyrene, phenanthrene, or anthracene; and some PAHs present in the environment accelerate the clearance of pyrene but reduce the level of 1-hydroxypyrene in urine (Lipniak-Gawlik 1998). Benz[a]anthracene, a weak carcinogen when applied simultaneously with dibenz[a,h]anthracene, inhibits the carcinogenic action of the latter in mouse skin; a similar case is made for benzo[e]pyrene or dibenz[a,c]anthracene applied to mouse skin shortly prior to initiation with 7,12-dimethylbenz[a]anthracene, or 3-methylcholanthrene (DiGiovanni and Slaga 1981a). Benzo[a]pyrene, a known carcinogen, interacts synergistically with cyclopenta[cd]pyrene, a moderately strong carcinogen found in automobile exhausts, according to results of mouse skin carcinogenicity studies (Rogan et al. 1983). Other PAH combinations were cocarcinogenic, such as benzo[e]pyrene, pyrene, and fluoranthene applied repeatedly with benzo[a]pyrene to the skins of mice (DiGiovanni and Slaga 1981a). Effective inhibitors of PAH-induced tumor development include selenium, Vitamin E, ascorbic acid, butylated hydroxytoluene, and hydroxyanisole (USEPA 1980). In addition, protective effects against PAH-induced tumor formation have been reported for various naturally occurring compounds such at flavones, retinoids, and Vitamin A (USEPA 1 80). Until these interaction effects are clarified, the results of single-substance laboratory studies may be extremely difficult to apply to field situations of suspected PAH contamination. Additional work is also needed on PAH dose-


response relationships, testing relevant environmental PAHs for carcinogenicity, and elucidating effects of PAH mixtures on tumor formation (Grimmer 1983).

25.5

RECOMMENDATIONS

No standards have been promulgated for PAHs by any regulatory agency for the protection of sensitive species of aquatic organisms or wildlife, although some criteria have been proposed for aquatic life protection (Table 25.9). This observation is not unexpected in view of several factors: 1. The paucity of data on PAH background concentrations in wildlife and other natural resources 2. The absence of information on results of chronic oral feeding studies of PAH mixtures and the lack of a representative PAH mixture for test purposes 3. The demonstrable — and as yet poorly understood — effects of biological modifiers, such as sex, age, and diet, and interaction effects of PAHs with inorganic and other organic compounds, including other PAHs.

Nevertheless, the growing database for aquatic life indicates a number of generalizations: (1) many PAHs are acutely toxic at concentrations between 50 and 1000 µg/L; (2) deleterious sublethal responses are sometimes observed at concentrations in the range of 0.1 to 5.0 µg/L; (3) uptake can be substantial, but depuration is usually rapid except in some species of invertebrates; and (4) wholebody burdens in excess of 300 µg benzo[a]pyrene/kg (and presumably other PAHs) in certain teleosts would be accompanied by a rise in the activity of detoxifying enzymes. Aquatic research has focused on PAHs because of their known relationship with carcinogenesis and mutagenesis (Black et al. 1988; Hawkins et al. 1990; Steard et al. 1990, 1991; Collier and Varanasi 1991; Smolowitz et al. 1992; Fernandez and L’Haridon 1994). Many reports exist of high incidences of cancer-like growths and developmental anomalies in natural populations of aquatic animals and plants, but none conclusively demonstrate the induction of cancer by exposure of aquatic animals to environmentally realistic levels of carcinogenic PAHs in the water column, diet, or sediments (Neff 1982b, 1985). However, studies by Baumann, Malins, Black, Varanasi, and their co-workers, among others, have now established that sediments heavily contaminated with PAHs from industrial sources were the direct cause of elevated PAH body burdens and elevated frequencies of liver neoplasia in fishes from these locales. Only a few sites containing high PAH concentrations in sediments have been identified (Couch and Harshbarger 1985), suggesting a need to identify and to evaluate other PAH-contaminated aquatic sites. Most fishery products consumed by upper trophic levels, including humans, contain PAH concentrations similar to those in green vegetables and smoked and charcoal-broiled meats, and would probably represent a minor source of PAH toxicity; however, consumption of aquatic organisms, especially filter-feeding bivalve molluscs, from regions severely contaminated with petroleum or PAH-containing industrial wastes, should be avoided (Jackim and Lake 1978; Neff 1982b). Neff (1982b) has suggested that repeated consumption of PAH-contaminated shellfish may pose a cancer risk to humans, and Al-Yakoob et al. (1994) aver that consumption of PAH-contaminated fish from the Arabian Gulf poses a risk to human health. If true, this needs to be evaluated using seabirds, pinnipeds, and other wildlife groups which feed extensively on molluscs and teleosts that are capable of accumulating high burdens of carcinogenic PAHs, in order to determine if similar risks exist. For avian wildlife, data are incomplete on PAH background concentrations and on acute and chronic toxicity. Studies with mallard embryos and PAHs applied to the egg surface showed toxic and adverse sublethal effects at concentrations between 0.036 and 0.18 µg PAH/kg whole egg (Hoffman and Gay 1981). Additional research is needed on petroleum-derived PAHs and their effects on developing embryos of seabirds and other waterfowl. There is an urgent need for specific avian biomarkers of PAH exposure (Murk et al. 1996).


PAH criteria for human health protection (Table 25.9) were derived from tests with small laboratory mammals, primarily rodents. Accordingly, these proposed criteria should become interim guidelines for protection of nonhuman mammalian resources pending acquisition of more definitive data. The proposed PAH criteria are controversial. Pucknat (1981) states that there was no way to quantify the potential human health risks incurred by the interaction of any PAH with other PAHs or with other agents in the environment, including tumor initiators, promoters, and inhibitors. The problem arises primarily from the diversity of test systems and bioassay conditions used for determining carcinogenic potential of individual PAHs in experimental animals, and is confounded by the lack of a representative PAH mixture for test purposes, the absence of data for animal and human chronic oral exposures to PAH mixtures, and the reliance on data derived from studies with benzo[a]pyrene to produce generalizations concerning environmental effects of PAHs — generalizations which may not be scientifically sound — according to Pucknat (1981). The USEPA (1980) emphasizes that only a small percentage of PAH compounds are known to be carcinogenic, and that measurements of total PAHs (i.e., the sum of all multiple fused-ring hydrocarbons having no heteroatoms) cannot be equated with carcinogenic potential. Furthermore, when the term “total PAHs” is used, the compounds being considered should be specified for each case. Lee and Grant (1981) state that an analysis of dose-response relationships for PAH-induced tumors in animals shows, in some cases, deviation from linearity in dose-response curves, especially at low doses, suggesting a two-stage model consistent with a linear nonthreshold pattern. Because overt tumor induction follows a dose-response relationship consistent with a multihit promotion process, the multihit component of carcinogenesis may be supplied by environmental stimuli not necessarily linked or related to PAH exposure. The well-documented existence of carcinogenic and anticarcinogenic agents strongly suggests that a time assessment of carcinogenic risk for a particular PAH can be evaluated only through a multifactorial analysis (Lee and Grant 1981). One of the most toxicologically significant processes involved in the response to PAH absorption is the interaction with drug metabolizing enzyme systems. The induction of this enzyme activity in various body tissues by PAHs and other xenobiotics is probably critical to the generation of reactive PAH metabolites at the target site for tumor induction. At present, wide variations occur in human and animal carcinogen-metabolizing capacity. Moreover, it has not yet been possible to definitely correlate enzyme activity with susceptibility to carcinogenesis. The obligatory coupling of metabolic activation with PAH-induced neoplasia in animals indicates that the modulation of drug-metabolizing enzymes is central to carcinogenesis (Lee and Grant 1981). PAHs from drinking water contribute only a small proportion of the average total human intake (Harrison et al. 1975). The drinking water quality criterion for carcinogenic PAH compounds is based on the assumption that each compound is as potent as benzo[a]pyrene and that the carcinogenic effect of the compounds is proportional to the sum of their concentrations (USEPA 1980). Based on an oral feeding study of benzo[a]pyrene in mice, the concentration of this compound estimated to result in additional risk of one additional case for every 100,000 individuals exposed (i.e, 10–5) is 0.028 µg/L. Therefore, with this assumption, the sum of the concentrations of all carcinogenic PAH compounds should be less than 0.028 µg/L in order to keep the lifetime cancer risk below 10–4 (USEPA 1980). The corresponding recommended criteria which may result in an incremental cancer risk of 10–6 and 10–7 over the lifetime are 0.0028 and 0.00028 µg/L, respectively (Table 25.9). If the above estimates are made for consumption of aquatic organisms only, the levels are 0.311 (10–5), 0.031(10–6), and 0.003(10–7) µg/kg, respectively (USEPA 1980). The use of contaminated water for irrigation can also spread PAHs into other vegetable foodstuffs (USEPA 1980). When vegetables grown in a PAH-polluted area are thoroughly washed and peeled, their contribution to total PAH intake in humans is not significant (Wang and Meresz 1982). Herbivorous wildlife, however, may ingest significant quantities of various PAHs from contaminated vegetables — but no data were available on this subject.


Table 25.9

Proposed PAH Criteria for Protection of Human Health and Aquatic Life

Criterion, PAH Group, and Units

Concentration

Referencei

HUMAN HEALTH Air Total PAHs, µg/m3 Total PAHs, daily intake, µgf Cyclohexane extractable fractions; coke oven emissions, coal tar products, µg/m3, 8–10 h-weighted average Benzene-soluble fractions; coal tar pitch volatiles, µg/m3, 8–h average Benzo[a]pyrene µg/m3 Daily intake, µgf Acceptable, µg/m3 AZ CT, IN FL, KS, MI ME NC PA TX Carcinogenic PAHs µg/m3 Daily intake, µgf

0.0109 0.164–0.251 100.0–150.0

1 1 1

200.0

1

0.0005 0.005–0.0115

1 1

0.79 for 1-h; 0.21 for 24 h; 0.00057 for 1 year 0.1 for 8-h 0.0003 for 1 year 0.0006 for 1 year 0.03 for 1 year 0.0007 for 1 year 0.03 for 30 min; 0.003 for 1 year

5 5 5 5 5 5 5

0.002 0.03–0.046

1 1

0.0135–0.2 0.027–0.4 4.0

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