Environmental Toxicants: Human Exposures and Their Health Effects, 3rd Edition

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ENVIRONMENTAL TOXICANTS

ENVIRONMENTAL TOXICANTS Human Exposures and Their Health Effects Third Edition

Edited by MORTON LIPPMANN

Copyright Ó 2009 by John Wiley & Sons, Inc. All rights reserved Published by John Wiley & Sons, Inc., Hoboken, New Jersey Published simultaneously in Canada No part of this publication may be reproduced, stored in a retrieval system, or transmitted in any form or by any means, electronic, mechanical, photocopying, recording, scanning, or otherwise, except as permitted under Section 107 or 108 of the 1976 United States Copyright Act, without either the prior written permission of the Publisher, or authorization through payment of the appropriate per-copy fee to the Copyright Clearance Center, Inc., 222 Rosewood Drive, Danvers, MA 01923, (978) 750-8400, fax (978) 750-4470, or on the web at www.copyright.com. Requests to the Publisher for permission should be addressed to the Permissions Department, John Wiley & Sons, Inc., 111 River Street, Hoboken, NJ 07030, (201) 748-6011, fax (201) 748-6008, or online at http://www.wiley.com/go/permission. Limit of Liability/Disclaimer of Warranty: While the publisher and author have used their best efforts in preparing this book, they make no representations or warranties with respect to the accuracy or completeness of the contents of this book and specifically disclaim any implied warranties of merchantability or fitness for a particular purpose. No warranty may be created or extended by sales representatives or written sales materials. The advice and strategies contained herein may not be suitable for your situation. You should consult with a professional where appropriate. Neither the publisher nor author shall be liable for any loss of profit or any other commercial damages, including but not limited to special, incidental, consequential, or other damages. For general information on our other products and services or for technical support, please contact our Customer Care Department within the United States at (800) 762-2974, outside the United States at (317) 572-3993 or fax (317) 572-4002. Wiley also publishes its books in a variety of electronic formats. Some content that appears in print may not be available in electronic formats. For more information about Wiley products, visit our web site at www.wiley.com. Library of Congress Cataloging-in-Publication Data: Environmental toxicants : human exposures and their health effects / [edited by] Morton Lippmann. – 3rd ed. p. ; cm. Includes bibliographical references and index. ISBN 978-0-471-79335-9 (cloth) 1. Environmental health. 2. Environmental technology. I. Lippmann, Morton. [DNLM: 1. Environmental Pollutants–adverse effects. 2. Environmental Exposure. 3. Environmental Health. 4. Environmental Pollutants–toxicity. WA 671 E615 2009] RA565.E58 2009 363.7–dc22 2008036266 Printed in the United States of America 10 9 8 7 6 5 4 3 2 1

CONTENTS

PREFACE CONTRIBUTORS

1 Introduction and Background 1.1 1.2 1.3 1.4 1.5 1.6 1.7 1.8 1.9 1.10 1.11 1.12 1.13

xvii

1

Characterization of Chemical Contaminants, 2 Human Exposures and Dosimetry, 7 Chemical Exposures and Dose to Target Tissues, 8 Concentration of Toxic Chemicals in Human Microenvironments, 9 Inhalation Exposures and Respiratory Tract Effects, 12 Ingestion Exposures and Gastrointestinal Tract Effects, 18 Skin Exposure and Dermal Effects, 19 Absorption through Membranes and Systemic Circulation, 20 Accumulation in Target Tissues and Dosimetric Models, 21 Indirect Measures of Past Exposures, 22 Characterization of Health, 23 Exposure–Response Relationships, 25 Study Options for Health Effects Studies, 31 References, 35

2 Perspectives on Individual and Community Risks 2.1 2.2 2.3 2.4

xv

39

Nature of Risk, 39 Identification and Quantification of Risks, 41 Risk Communication, 46 Risk Reduction, 49 References, 52 v

vi

CONTENTS

3 Reducing Risks—An Environmental Engineering Perspective 3.1 3.2 3.3 3.4 3.5 3.6

Introduction, 55 Environmental Risk-Based Decision Making, 56 Applications and Use, 60 Recent Information, 67 Integrated Assessments, 71 Summary, 72 References, 73

4 Clinical Perspective on Respiratory Toxicology 4.1 4.2 4.3 4.4 4.5 4.6 4.7

5.2 5.3 5.4 5.5 5.6

107

The Life Cycle of a Chemical: Many Points for Possible Intervention, 108 The Knowledge Base for the Identification of Hazard Control Strategies, 109 Industrial Hygiene and Occupational Health Programs: Implementing the Knowledge Base, 111 Product Stewardship, 114 Responsible CareÒ, 117 Concluding Perspective, 119

6 Drinking Water Disinfection By-Products 6.1 6.2 6.3

77

Concepts of Exposure, 78 Tools for Studying Individuals, 79 Tools for Studying Populations, 88 Cardiovascular Responses, 95 Limitations of Clinical and Epidemiological Assessments of the Effects of Inhaled Agents, 96 Advice and Counseling of Patients, 97 Summary, 99 References, 100

5 Industrial Perspectives: Translating the Knowledge Base into Corporate Policies, Programs, and Practices for Health Protection 5.1

55

Introduction, 121 Chemical Methods of Disinfection, 122 Chemical Nature and Occurrence of Disinfectant By-Products, 124 6.4 Associations of Human Disease with Drinking Water Disinfection, 132 6.5 General Toxicological Properties of Disinfectants, 144 6.6 General Toxicological Properties of Disinfectant By-Products, 145 6.7 Carcinogenic Properties of Disinfectants, 154 6.8 Carcinogenic By-Products of Disinfectants, 154 6.9 Effects of Disinfectants and Their By-Products on Reproduction, 165 6.10 Effects on Development, 168 6.11 By-Products of Potential Interest, 170

121

CONTENTS

vii

6.12 Summary and Conclusions, 172 Glossary, 174 References, 176 7 Food 7.1 7.2 7.3 7.4 7.5 7.6 7.7 7.8

197 Introduction, 197 Legal and Regulatory Framework in the United States, 201 Toxicity Test Requirements and Safety Criteria, 203 Substances Intentionally Added to Food, 208 Food Contaminants of Industrial Origin, 216 Constituents and Contaminants of Natural Origin, 219 Food Safety in the European Union, 229 Summary and Conclusion, 234 Acronyms, 235 References, 235

8 Volatile Organic Compounds and Sick Building Syndrome 8.1 8.2 8.3 8.4 8.5 8.6

Introduction, 241 Prevalence of Exposures to Volatile Organic Compounds, 242 Health and Volatile Organic Compounds, 245 Prevalence of the Sick Building Syndrome, 247 Dose–Response Relationships for Health Effects Caused by Low-Level VOC Exposure, 249 Guidelines for Volatile Organic Compounds in Nonindustrial Indoor Environments-Principles for Establishment of Guidelines, 251 References, 254

9 Formaldehyde and Other Aldehydes 9.1 9.2 9.3

257

Background, 257 Single-Exposure Health Effects, 269 Effects of Multiple Exposures, 281 References, 292

10 Ambient Air Particulate Matter 10.1 10.2 10.3 10.4

241

Sources and Pathways for Human Exposure, 318 Ambient Air PM Concentrations, 323 Extent of Population Exposures to Ambient Air PM, 326 Nature of the Evidence for Human Health Effects of Ambient Air PM, 328 10.5 Epidemiological Evidence for Human Health Effects of Ambient Air PM, 329 10.6 Discussion and Current Knowledge on the Health Effects of PM, 354

317

viii

CONTENTS

10.7 Standards and Exposure Guidelines, 356 References, 359 11 Arsenic

367

11.1 Introduction, 367 11.2 Physical and Chemical Properties of Environmental as and Its Compounds, 368 11.3 Environmental Exposures to the General Population: Sources and Standards, 371 11.4 Pathways and Kinetics for in vivo Uptake, Distribution, and Elimination, 374 11.5 As Essentiality, 375 11.6 Health Effects and Exposure–Response Relationships, 376 11.7 Biomarkers of Exposure, Susceptibility, and Effect, 380 11.8 Mitigating Effects and Controlling Exposures, 381 References, 383 12 Asbestos and Other Mineral and Vitreous Fibers

395

12.1 12.2 12.3 12.4 12.5 12.6 12.7

Important Special Properties of Fibers, 395 Exposures to Fibers, 399 Fiber Deposition in the Respiratory Tract, 402 Fiber Retention, Translocation, Disintegration, and Dissolution, 404 Properties of Fibers Relevant to Disease, 413 Fiber-Related Diseases/Processes, 413 Review of Biological Effects of Size-Classified Fibers in Animals and Humans, 415 12.8 Critical Fiber Parameters Affecting Disease Pathogenesis, 420 12.9 Exposure–Response Relationships for Asbestos-Related Lung Cancer and Mesothelioma: Human Experience, 429 12.10 Risk Assessment Issues, 438 12.11 Key Factors Affecting Fiber Dosimetry and Toxicity: Recapitulation and Synthesis, 443 Acknowledgments, 446 Acronyms, 446 References, 446 13 Benzene 13.1 13.2 13.3 13.4 13.5

Benzene Exposure, 460 Uptake, 462 Metabolism and Disposition, 462 Mechanisms of Toxicity, 471 Risk Assessment, 482 References, 486

14 Carbon Monoxide 14.1 14.2

459

Introduction, 499 CO Exposure and Dosimetry, 500

499

CONTENTS

14.3 14.4 14.5 14.6 14.7

Mechanisms of CO Toxicity, 501 Populations at Risk of Health Effects Due to CO Exposure, 502 Regulatory Background, 503 Health Effects of CO, 505 Summary and Conclusions, 515 Acknowledgments, 517 References, 517

15 Chromium 15.1 15.2 15.3 15.4 15.5

18.3 18.4 18.5 18.6 18.7

633

Introduction, 633 Sources, 634 Toxicological Effects and Mechanisms of Action, 640 Mechanisms of Action, 643 References, 651

18 Endocrine Active Chemicals: Broadening the Scope 18.1 18.2

551

Historical Overview, 551 Composition of Diesel Exhaust, 553 Exposures to Diesel Exhaust, 559 Health Effects, 561 Current Issues, 609 Acknowledgments, 613 References, 613

17 Dioxins and Dioxin-Like Chemicals 17.1 17.2 17.3 17.4

529

Introduction, 529 Essentiality, 529 Environmental Exposures, 530 Toxicological Effects, 535 Exposure Guidelines and Standards, 543 References, 544

16 Diesel Exhaust 16.1 16.2 16.3 16.4 16.5

ix

Introduction, 661 Biomarkers: Terminology from Various Disciplines, 664 End Points and Clinical Signs Associated with Endocrine Activity, 666 Environmental Chemicals and End Points: Case Examples, 675 Developmental Origins of Health and Disease, 681 Transgenerational Effects, 684 Conclusion, 686 References, 687

661

x

CONTENTS

19 Secondhand Smoke 19.1 19.2 19.3 19.4 19.5 19.6

703

Exposure to Secondhand Smoke, 705 Health Effects of Involuntary Smoking in Children, 711 Health Effects of Involuntary Smoking in Adults, 722 SHS and Coronary Heart Disease, 730 Respiratory Symptoms and Illnesses in Adults, 734 Summary, 740 References, 741

20 Lead and Compounds

757

20.1 20.2

Introduction, 757 Physical/Chemical Properties and Behavior of Lead and Its Compounds, 758 20.3 Lead in the Environment and Human Exposure, 761 20.4 Lead Absorption, 766 20.5 Distribution, 771 20.6 Kinetics, 774 20.7 Biomarkers, 781 20.8 Health Effects, 785 20.9 Mechanisms Underlying Lead Toxicity, 792 20.10 Treatment of Lead Toxicity, 796 References, 798 21 Mercury 21.1 21.2 21.3 21.4 21.5 21.6 21.7 21.8

Introduction, 811 Chemistry, 811 Sources, 812 Environmental Exposures, 813 Occupational Exposures, 815 Kinetics and Metabolism, 816 Health Effects, 818 Prevention, 820 References, 821

22 Nitrogen Oxides 22.1 22.2 22.3 22.4 22.5 22.6

823

Introduction, 823 Sources, 823 Nitrogen Dioxide, 824 Nitric Oxide, 845 Nitric/Nitrous Acid, 848 Inorganic Nitrates, 849 References, 851

23 Ozone 23.1 23.2

811

869 Introduction, 869 Background on Exposures and Health-Related Effects, 873

CONTENTS

xi

23.3 23.4 23.5 23.6 23.7

Effects of Short-Term Exposures to Ozone in Humans, 877 Factors Affecting the Variability of Responsiveness in Humans, 890 Studies of Populations Exposed to Ozone in Ambient Air, 892 Effects Observed in Studies in Laboratory Animals, 900 Determinants of Responsiveness to Ozone Exposures in Animal Studies, 901 23.8 Effects of Multiple Day and Ambient Episode Exposures, 908 23.9 Chronic Effects of Ambient Ozone Exposures, 910 23.10 Ambient Air Quality Standards and Guidelines, 917 23.11 Summary and Conclusions, 920 Acknowledgment, 922 References, 922 24 Pesticides 24.1 24.2 24.3 24.4 24.5 24.6 24.7 24.8 24.9

Evolving Patterns of Pesticide Use, 938 Export of Hazardous Pesticides, 939 Exposure to Pesticides, 939 Epidemiology of Acute Pesticide Poisoning, 942 Toxicity of Pesticides, 943 Pesticides and Endocrine/Reproductive Toxicity, 949 Pesticides and Childhood Cancer, 950 Legislative Framework, 950 Conclusion: Issues for the Future, 952 References, 953

25 Sulfur Oxides—SO2, H2SO4, NH4HSO4, and (NH4)2SO4 25.1 25.2 25.3

1001

Background, 1003 Philosophical Approaches, 1004 Standards Development, 1005 Current Developments, 1010 Protective Measures, 1012 Conclusions, 1014 Glossary, 1015 References, 1016

27 Sources, Levels and Effects of Manmade Ionizing Radiation and Radioactivity 27.1 27.2

957

Sources and Exposures, 957 Health Effects, 961 Ambient Air Quality Standard and Guidelines, 989 Acknowledgments, 991 References, 991

26 Microwaves and Electromagnetic Fields 26.1 26.2 26.3 26.4 26.5 26.6 26.7

937

Source Documents, 1021 Special Units, 1022

1021

xii

CONTENTS

27.3 27.4 27.5 27.6 27.7 27.8 27.9 27.10 27.11

Sources of Manmade Radioactivity and Radiation, 1024 Nuclear Fuel Cycle, 1025 Discussion of Radiation Doses from the Nuclear Fuel Cycle, 1038 Nuclear Weapons Complex, 1043 Local, Tropospheric, and Global Fallout, 1048 Medical Exposures, 1050 Industrial Uses (Other than the Nuclear Fuel Cycle), 1054 Consumer Products, 1055 Overview of Potential Health Impacts of Natural and Manmade Sources of Radioactivity, 1057 References, 1066

28 Noise: Its Effects and Control 28.1 28.2 28.3 28.4 28.5 28.6 28.7 28.8 28.9 28.10 28.11 28.12

Definitions of Sound and Noise, 1071 Noise Exposure is Widespread and Annoying, 1072 Effects of Loud Sounds and Noise on Hearing, 1075 Noise as a Stressor, 1076 Noise and Sleep Interference, 1077 Noise and Mental Health, 1077 Noise Affects Children’s Cognitive, Language and Learning Skills, 1078 Impacts of Low-Frequency Noise, 1079 Civility, Responsibility, and Noise, 1079 Controlling Noise, 1080 Education and Public Awareness, 1084 Summary, 1084 References, 1085

29 Radon and Lung Cancer 29.1 29.2 29.3 29.4 29.5 29.6 29.7 29.8 29.9 29.10 29.11 29.12

1089

Radon and Lung Cancer, 1089 Outdoor Radon, 1093 Indoor Radon, 1097 The Other Radon, 220Rn, Thoron, 1100 Radon Epidemiology in Underground Mines, 1100 Residential Epidemiology, 1102 Lung Dosimetry, 1104 Lung Cancer Models for Humans, 1107 Childhood Exposure, 1113 Animal Studies, 1114 Smoking and Radon, 1114 Summary, 1115 References, 1116

30 Ultraviolet Radiation 30.1 30.2 30.3

1071

Introduction, 1121 Pathways for Human Exposure, 1122 Sources of Ultraviolet Radiation, 1124

1121

CONTENTS

30.4 30.5 30.6 30.7 30.8 30.9 30.10 30.11 30.12 30.13 30.14

INDEX

xiii

Biological Mechanisms Leading to Health Effects, 1135 Ocular Effects, 1135 Nonmalignant Skin Effects, 1138 Skin Cancer, 1140 Malignant Melanoma, 1142 Immune System Effects, 1146 Populations at Special Risk: Ocular Damage, 1147 Populations at Special Risks: Skin Effects, 1148 Applicable Standards and Exposure Guidelines, 1150 Techniques for Evaluating Actual or Potential Exposures, 1152 Summary, 1156 References, 1157 1163

PREFACE

This is the third edition of Environmental Toxicants: Human Exposures and Their Health Effects. It provides updated versions of chapters that appeared in the first (1992) and second (2000) editions, and it broadens the coverage to include two new toxicant categories (one is arsenic and its compounds, and the other is endocrine disrupting chemicals). As before, it is focused on providing current knowledge on environmental health challenges to people in our homes and communities resulting from exposures to chemical and physical agents that they encounter in the course of their daily lives. This book remains unique in terms of its depth of coverage on a limited number of environmental agents that are known to have, or are highly likely to have, adverse health effects following exposures that are within the ranges that occur in contemporary populations in economically developed countries. Extrapolation of likely effects in developing countries, where toxicant exposures may be substantially higher, need to be made with caution, since susceptibility to adverse effects may differ as a result of differences in diet, pre-existing diseases, thermal stresses, and access to modern health care. Chapter 1 has been expanded to include discussions of study options for increasing our knowledge of biological responses to environmental toxicant exposures, as well as of new and developing methods for the elucidation of responses at the molecular level. I gratefully acknowledge the outstanding contributions of the other chapter authors who are my colleagues and peers. They are all outstanding and widely recognized professionals with many demands on their time, and this unique book would not have been possible without their generous commitment. Periodic revisions of the content of the chapters herein are necessary because of our everincreasing knowledge base, which has been facilitated by the development of new and improved measurement and modeling, and data management technologies. These technologies, and the growth of interdisciplinary investigations of complex phenomena, have enabled investigatory teams to go beyond the identification of statistically significant associations between environmental exposures and health-related responses in human populations, laboratory animal cohorts, and cell cultures in vitro, to the underlying biological pathways and mechanisms that are applicable to realistic exposure levels. While xv

xvi

PREFACE

notable progress has been made in environmental health sciences in recent years, significant challenges remain, not the least of which is access to research funding from government and private sources at a time when our collective capacities are increasing for characterizing (1) exposures and their geographic and temporal distributions; (2) biological mechanisms responsible for the adverse effects produced by environmental exposures; (3) susceptibility factors that account for the generally large interindividual variability in responses to exposure; and (4) exposure–response relationships for sensitive population segments. Another challenge is that the populations of both the general public and the environmental health research community are aging. Older people are clearly a susceptible population to many environmental toxicants, and the research needed to identify means of recognizing, evaluating, and controlling exposures to these toxicants will require both additional research funding and recruitment and training of young investigators who can carry out such research over at least several decades into the future. Some of these new trainees may well be the authors of chapters in future editions of this reference volume. I hereby recognize the contributions of those who, in addition to writing chapters, made substantial contributions to the completion of this edition. In particular, I want to recognize Toni Moore, Anita Parkhurst, and Angela Muniz for their diligent and effective management of the text preparation and presentation, and Gordon Cook for the preparation of many of the figures. Finally, my own contributions would not have been possible without the cooperation and patience of my wife, Janet. New York University School of Medicine

MORTON LIPPMANN

CONTRIBUTORS

Donald R. Bergfelt, Division of Exposure Assessment, Office of Science Coordination and Policy, Office of Prevention, Pesticides and Toxic Substances, United States Environmental Protection Agency, 1200 Pennsylvania Avenue, NW, Washington, DC 20460, USA Arline L. Bronzaft, 505 East 79th Street, 8B, New York, NY 10021, USA Richard J. Bull, MoBull Consulting, 1928 Meadows Drive North, Richland, WA 99352, USA James S. Bus, Toxicology Research Laboratory, Dow Chemical Co., 1803 Building, Midland, MI 48674, USA Luz Claudio, Department of Community and Preventive Medicine, Mount Sinai School of Medicine, New York, NY 10029, USA Mitchell D. Cohen, Nelson Institute of Environmental Medicine, New York University School of Medicine, 57 Old Forge Road, Tuxedo, NY 10987, USA Norman Cohen, New York University (Retired) Francis Colville, Radiofrequency Program, U.S. Army Center for Health Promotion and Preventive Medicine, Aberdeen Proving Ground, MD, 21010-5403, USA Nigel Cridland, National Radiological Protection Board, Chilton, Didcot, Oxon OX11 0RQ, UK Colin Driscoll, National Radiological Protection Board (Retired) Michael A. Gallo, Environmental & Occupational Health Science Institute, 681 Frelinghuysen Road, Piscataway, NJ 08855-1179, USA Eric Garshick, Pulmonary and Critical Care Medicine Section, VA Boston Healthcare System; Channing Laboratory, Brigham and Women’s Hospital; and Harvard Medical School, Boston, MA, USA xvii

xviii

CONTRIBUTORS

Bernard D. Goldstein, Graduate School of Public Health, University of Pittsburgh, Pittsburgh, PA, 15261, USA Philippe Grandjean, Department of Environmental Medicine, Odense University, Winslowparken 17, DK-5000 Odense, Denmark Lester D. Grant, 517 Colony Woods Drive, Chapel Hill, NC, USA K. Christiana Grim, Division of Exposure Assessment, Office of Science Coordination and Policy, Office of Prevention, Pesticides and Toxic Substances, United States Environmental Protection Agency, 1200 Pennsylvania Avenue, NW, Washington, DC 20460, USA Naomi H. Harley, Department of Environmental Medicine, New York University Medical School, 550 First Avenue, New York, NY 10016, USA Fred D. Hoerger, Dow Chemical (Retired) Michael T. Kleinman, Department of Community and Environmental Medicine, University of California at Irvine, Irvine, CA 92697-1825, USA Philip J. Landrigan, Department of Community Medicine, Mount Sinai Medical Center, Box 1057, New York, NY 10029-6574, USA George D. Leikauf, Department of Environmental and Occupational Health, Graduate School of Public Health, Bridgestone Point Bldg, Suite 359, 100 Technology Drive, Pittsburgh, PA 15219, USA Morton Lippmann, New York University School of Medicine, 21 Old Forge Lane, Tarrytown, NY 10591, USA Raymond C. Loehr, 19360 Magnolia Grove Square No. 405, Lansdowne, VA 20176, USA Kathryn R. Mahaffey, Office of Science Coordination and Policy, Office of Prevention, Pesticides and Toxic Substances, United States Environmental Protection Agency, 1200 Pennsylvania Avenue, NW, Washington, DC 20460, USA Joe L. Mauderly, Inhalation Toxicology Research Institute, P.O. Box 5890, Albuquerque, NM 87185, USA John J. Mauro, 209 Ueland Road, Red Bank, NJ 07701, USA Jessica C. Meiller, Division of Exposure Assessment, Office of Science Coordination and Policy, Office of Prevention, Pesticides and Toxic Substances, United States Environmental Protection Agency, 1200 Pennsylvania Avenue, NW, Washington, DC 20460, USA Lars Mølhave, Institute of Environmental and Occupational Medicine, University of Aarhus, DK-8000 Aarhus C, Denmark Gila I. Neta, Department of Epidemiology, Bloomberg School of Public Health, Johns Hopkins University, Baltimore, MD, USA Jesper B. Nielsen, Institute of Public Health, University of Southern Denmark, Winslowparken 17, DK-5000 Odense, Denmark

CONTRIBUTORS

xix

Larry W. Rampy, Dow Chemical (Retired) Douglas A. Rausch, Dow Chemical (Retired) Joseph V. Rodricks, The Life Sciences Consultancy LLC, 750 17th Street, NW, Suite 1000, Washington, DC 20006, USA Toby G. Rossman, Nelson Institute of Environmental Medicine, New York University School of Medicine, 57 Old Forge Road, Tuxedo, NY 10987, USA Jonathan M. Samet, Department of Epidemiology, The Johns Hopkins University, Suite 6039, 615 N. Wolfe Street, Baltimore, MD 21205-2179, USA Richard B. Schlesinger, Dyson College of Arts & Sciences, Pace University, 861 Bedford Road, Pleasantville, NY 10570, USA David H. Sliney, 406 Streamside Drive, Fallston, MD 21047-2806, USA Shirlee W. Tan, The Smithsonian Institution, National Zoological Park, 3000 Connecticut Avenue, NW, Washington, DC, USA Arthur C. Upton, 250 East Alameda, Apartment 636, Santa Fe, NM 87501, USA Mark J. Utell, Pulmonary Unit, University of Rochester Medical Center, Box 692, Rochester, NY 14642-8692, USA Sophia S. Wang, Division of Cancer Epidemiology and Genetics, National Cancer Institute, Washington, DC, USA Gisela Witz, Department of Environmental and Occupational Medicine, Robert Wood Johnson Medical School, Piscataway, NJ, USA

1 INTRODUCTION AND BACKGROUND Morton Lippmann and George D. Leikauf

This book identifies and critically reviews current knowledge on human exposure to selected chemical agents and physical factors in the ambient environment and the effects of such exposures on human health. It provides a state-of-the-art knowledge base essential for risk assessment for exposed individuals and populations to guide public health authorities, primary care physicians, and industrial managers having to deal with the consequences of environmental exposure. Aside from professionals in public health, medicine, and industry who may use this book to guide their management functions, the volume can also be used in graduate and postdoctoral training programs in universities and by toxicologists, clinicians, and epidemiologists in research as a resource for the preparation of research proposals and scientific papers. The subject is environmental toxicants, that is, chemical or physical agents released into the general environment that can produce adverse health effects among large numbers of people. Such effects are usually subclinical, except when cumulative changes lead to chronic effects after long exposure. Short-term responses following acute exposures are often manifest as transient alterations in physiological function that may, in some sensitive members of the population, be of sufficient magnitude to be considered adverse. Each of the specific topic chapters has a thorough discussion of the extent of human exposure as well as of toxic responses. The four chapters on the uses of the data for risk assessment, risk management, clinical applications, and industrial operations provide guidance for those performing individual and/or collective population hazard evaluations. The first provides individuals and public agency personnel with a basis for decisions on risk avoidance and relative risk assessment. The second outlines the operational philosophies and techniques used by environmental engineers in scoping and managing environmental risks. The third enables the primary care physician to recognize diseases and symptoms associated with exposures to environmental toxicants and to provide counsel to patients. The fourth assists decision makers in industry in evaluating the potential impacts of their plant operations and products on public health.

Environmental Toxicants, Third Edition Edited by Morton Lippmann Copyright 2009 John Wiley & Sons, Inc.

1

2

INTRODUCTION AND BACKGROUND

Although many books provide brief reviews of hundreds of chemicals encountered in the work environment at levels that can cause demonstrable health effects, both acute and chronic, they contain relatively little information on the effects of low-level exposures on large populations of primary interest in environmental health and risk assessment. This book has been designed to provide in-depth, critical reviews of the environmental toxicants of contemporary public health concern.

1.1 CHARACTERIZATION OF CHEMICAL CONTAMINANTS 1.1.1

Concentration Units

In environmental science, confusion often arises from the use of the same or similar sounding terms having different meanings in different contexts. This is especially true in describing the concentrations of air and water contaminants. Solutes are frequently expressed in parts per million (ppm) or parts per billion (ppb). However, when used for air contaminants, the units are molar or volume fractions, whereas when used for water contaminants, they are weight fractions. This problem can be avoided by expressing all fluid contaminant concentrations as the weight of contaminant per unit volume (e.g., m3 or L) of fluid. In air, the units generally used are mg/m3 or mg/m3, whereas in water they are most often mg/L or mg/L. 1.1.2

Air Contaminants

Chemical contaminants can be dispersed in air at normal ambient temperatures and pressures in gaseous, liquid, and solid forms. The latter two represent suspensions of particles in air and were given the generic term “aerosols” by Gibbs (1924) on the basis of analogy to the term “hydrosol,” used to describe disperse systems in water. On the contrary, gases and vapors, which are present as discrete molecules, form true solutions in air. Particles consisting of moderate- to high-vapor-pressure materials tend to evaporate rapidly, since those small enough to remain suspended in air for more than a few minutes (i.e., those smaller than about 10 mm) have large surface-to-volume ratios. Some materials with relatively low vapor pressures can have appreciable fractions in both the vapor and aerosol forms simultaneously. 1.1.2.1 Gases and Vapors Once dispersed in air, contaminant gases and vapors generally form mixtures so dilute that their physical properties, such as density, viscosity, enthalpy, and so on are indistinguishable from those of clean air. Such mixtures may be considered to follow ideal gas law relationships. There is no practical difference between a gas and a vapor except that the latter is generally considered to be the gaseous phase of a substance that is normally a solid or liquid at room temperature. While dispersed in the air, all molecules of a given compound are essentially equivalent in their size and probabilities of contact with ambient surfaces, respiratory tract surfaces, and contaminant collectors or samplers. 1.1.2.2 Aerosols Aerosols, being dispersions of solid or liquid particles in air, have the very significant additional variable of particle size. Size affects particle motion and, hence, the probabilities for physical phenomena such as coagulation, dispersion, sedimentation, impaction onto surfaces, interfacial phenomena, and light-scattering properties. It is not possible to fully characterize a given particle by a single size parameter. For example, a

CHARACTERIZATION OF CHEMICAL CONTAMINANTS

(a)

(b)

(c)

20

Frequency

Mean

Particle size (µm)

Median

Frequency

40 84.1% Value

Mode

0

3

20 10 Particle size (µm)

1

10 100 Particle size (µm)

10

50% Value

5

1

1

10 50 90 99 Cumulative (%)

FIGURE 1.1 Particle size distribution data. (a) Plotted on linear coordinates. (b) Plotted on a logarithmic size scale. (c) In practice, logarithmic probability coordinates are used to display the percentage of particles less than a specific size versus that size. The geometric standard deviation (sg) of the distribution is equal to the 84.1% size/50% size.

particle’s aerodynamic properties depend on density and shape as well as linear dimensions, and the effective size for light scattering is dependent on refractive index and shape. In some special cases, all of the particles are essentially the same in size. Such aerosols are considered to be monodisperse. Examples are natural pollens and some laboratorygenerated aerosols. More typically, aerosols are composed of particles of many different sizes and hence are called heterodisperse or polydisperse. Different aerosols have different degrees of size dispersion. It is, therefore, necessary to specify at least two parameters in characterizing aerosol size: a measure of central tendency, such as a mean or median, and a measure of dispersion, such as an arithmetic or geometric standard deviation. Particles generated by a single source or process generally have diameters following a lognormal distribution; that is, the logarithms of their individual diameters have a Gaussian distribution. In this case, the measure of dispersion is the geometric standard deviation, which is the ratio of the 84.16 percentile size to the 50th percentile size (Fig. 1.1). When more than one source of particles is significant, the resulting mixed aerosol will usually not follow a single lognormal distribution, and it may be necessary to describe it by the sum of several distributions. 1.1.3

Particle Characteristics

There are many properties of particles, other than their linear size, that can greatly influence their airborne behavior and their effects on the environment and health. These include Surface: For spherical particles, the surface varies as the square of the diameter. However, for an aerosol of given mass concentration, the total aerosol surface increases with decreasing particle size. Airborne particles have much greater ratios of external surface to volume than do bulk materials and, therefore, the particles can dissolve or participate in surface reactions to a much greater extent than would massive samples of the same materials. Furthermore, for nonspherical solid particles

4


or aggregate particles, the ratio of surface to volume is increased, and for particles with internal cracks or pores, the internal surface area can be much greater than the external area. Volume: Particle volume varies as the cube of diameter; therefore, the few largest particles in an aerosol tend to dominate its volume concentration. Shape: A particle’s shape affects its aerodynamic drag as well as its surface area and therefore its motion and deposition probabilities. Density: A particle’s velocity in response to gravitational or inertial forces increases as the square root of its density. Aerodynamic diameter: The diameter of a unit-density sphere having the same terminal settling velocity as the particle under consideration is equal to its aerodynamic diameter. Terminal settling velocity is the equilibrium velocity of a particle that is falling under the influence of gravity and fluid resistance. Aerodynamic diameter is determined by the actual particle size, the particle density, and an aerodynamic shape factor. 1.1.3.1 Types of Aerosols Aerosols are generally classified in terms of their processes of formation. Although the following classification is neither precise nor comprehensive, it is commonly used and accepted in the industrial hygiene and air pollution fields: Dust: An aerosol formed by mechanical subdivision of bulk material into airborne fines having the same chemical composition. A general term for the process of mechanical subdivision is comminution, and it occurs in operations such as abrasion, crushing, grinding, drilling, and blasting. Dust particles are generally solid and irregular in shape and have diameters greater than 1 mm. Fume: An aerosol of solid particles formed by condensation of vapors formed at elevated temperatures by combustion or sublimation. The primary particles are generally very small (less than 0.1 mm) and have spherical or characteristic crystalline shapes. They may be chemically identical to the parent material, or they may be composed of an oxidation product such as a metal oxide. Since they may be formed in high number concentration, they often rapidly coagulate, forming aggregate clusters of low overall density. Smoke: An aerosol formed by condensation of combustion products, generally of organic materials. The particles are generally liquid droplets with diameters of less than 0.5 mm. Mist: A droplet aerosol formed by mechanical shearing of a bulk liquid, for example, by atomization, nebulization, bubbling, or spraying. The initial droplet size can cover a very large range, usually from about 2 mm to greater than 50 mm. Fog: An aqueous aerosol formed by condensation of water vapor on atmospheric nuclei at high relative humidities. The droplet sizes are generally greater than 1 mm. Smog: A popular term for a pollution aerosol derived from a combination of smoke and fog. It is now commonly used for any atmospheric pollution mixture. Haze: A submicrometer-sized aerosol of hygroscopic particles that take up water vapor at relatively low relative humidities. Aitken or condensation nuclei (CN): Very small atmospheric particles (mostly smaller than 0.1 mm) formed by combustion processes and by chemical conversion from gaseous precursors.

CHARACTERIZATION OF CHEMICAL CONTAMINANTS

5

Accumulation mode: A term given to the particles in the ambient atmosphere ranging from 0.1 to about 1.0 mm, and extending up to 2.5 mm for hygroscopic particles in humid atmospheres. These particles generally are spherical, have liquid surfaces, and form by coagulation and condensation of smaller particles that derive from gaseous precursors. Being too few for rapid coagulation, and too small for effective sedimentation, they tend to accumulate in the ambient air. Coarse particle mode: Ambient air particles larger than about 2.5 mm and generally formed by mechanical processes and surface dust resuspension. 1.1.3.2 Aerosol Characteristics Aerosols have integral properties that depend upon the concentration and size distribution of the particles. In mathematical terms, these properties can be expressed in terms of certain constants or “moments” of the size distribution (Friedlander, 1977). Some integral properties such as light-scattering ability or electrical charge depend on other particle parameters as well. Some of the important integral properties are: Number concentration: The total number of airborne particles per unit volume of air, without distinction as to their sizes, is the zeroth moment of the size distribution. In current practice, instruments are available that count the numbers of particles of all sizes from about 0.005 to 50 mm. In many specific applications, such as fiber counting for airborne asbestos, a more restricted size range is specified. Surface concentration: The total external surface area of all the particles in the aerosol, which is the second moment of the size distribution, may be of interest when surface catalysis or gas adsorption processes are of concern. Aerosol surface is one factor affecting light-scatter and atmospheric-visibility reductions. Volume concentration: The total volume of all the particles, which is the third moment of the size distribution, is of little intrinsic interest in itself. However, it is closely related to the mass concentration, which for many environmental effects is the primary parameter of interest. Mass concentration: The total mass of all the particles in the aerosol is frequently of interest. The mass of a particle is the product of its volume and density. If all of the particles have the same density, the total mass concentration is simply the volume concentration times the density. In some cases, such as “respirable,” “thoracic,” and “inhalable” dust sampling (Vincent, 1999), the parameter of interest is the mass concentration over a restricted range of particle size. In these applications, particles outside the size range of interest are excluded from the integral. Dustfall: The mass of particles depositing from an aerosol onto a unit surface per unit time is proportional to the fifth moment of the size distribution. Dustfall has long been of interest in air pollution control because it provides an indication of the soiling properties of the aerosol. Light scatter: The ability of airborne particles to scatter light and cause a visibility reduction is well known. Total light scatter can be determined by integrating the aerosol surface distribution with the appropriate scattering coefficients. 1.1.4

Water Contaminants

Chemical contaminants can be found in water, in solution, or as hydrosols; the latter are immiscible solid or liquid particles in suspension. An aqueous suspension in liquid particles

6


is generally called an emulsion. Many materials with relatively low aqueous solubility will be found in both dissolved and suspended forms. 1.1.4.1 Dissolved Contaminants Water is known as the universal solvent. Although there are many compounds that are not completely soluble in water, there are a few that do not have some measurable solubility. In fact, the number of chemical contaminants in natural waters is primarily a function of the sensitivity of the analyses. For organic compounds in rivers and lakes, it has been observed that as the limits of detection decrease by an order of magnitude, the numbers of compounds detected increase by an order of magnitude, so that one might expect to find at least 10–12 g/L (approximately 1010 molecules/L) of each of the million organic compounds reported in the literature (NIEHS, 1977). Similar considerations undoubtedly apply to inorganic chemicals as well. 1.1.4.2 Dissolved Solids Water-quality criteria generally include a nonspecific parameter called “dissolved solids.” However, it is customary to exclude natural mineral salts such as sodium chloride from this classification. Also, water criteria for specific toxic chemicals dissolved in water are frequently exceeded without there being an excessive total dissolved-solids content. 1.1.4.3 Dissolved Gases Compounds dissolved in water may also exist in the gaseous phase at normal temperatures and pressures. Some of these, such as hydrogen sulfide (HS2), and ammonia (NH3), which are generated by decay processes, are toxicants. Oxygen (O2) is the most critical of the dissolved gases with respect to water quality. It is essential to most higher aquatic life forms and is needed for the oxidation of most of the organic chemical contaminants to more innocuous forms. Thus, a critical parameter of water quality is the concentration of dissolved oxygen (DO). Another important parameter is the extent of the oxygen “demand” associated with contaminants in the water. The most commonly used index of oxygen demand is the 5-day biochemical oxygen demand (BOD after 5 days of incubation). Another is the chemical oxygen demand (COD). 1.1.4.4 Suspended Particles A nonspecific water-quality parameter that is widely used is “suspended solids.” The stability of aqueous suspensions depends on particle size, density, and charge distributions. The fate of suspended particles depends on a number of factors, and particles can dissolve, grow, coagulate, or be ingested by various life forms in the water. They can become “floating solids” or part of an oil film, or they can fall to the bottom to become part of the sediments. There are many kinds of suspended particles in natural waters, and not all of them are contaminants. Any moving water will have currents that cause bottom sediments to become resuspended. Also, natural runoff will carry soil and organic debris into lakes and streams. In any industrialized area, such sediment and surface debris will always contain some chemicals considered to be contaminants. However, a large proportion of the mass of such suspended solids would usually be “natural,” and would not be considered as contaminants. The suspended particles can have densities that are less than, equal to, or greater than that of the water, so that the particles can rise as well as fall. Furthermore, the effective density of particles can be reduced by the attachment of gas bubbles. Gas bubbles form in water when the water becomes saturated and cannot hold any more of the gas in solution. The solubility of gases in water varies inversely with temperature. For example, oxygen saturation of fresh water is 14.2 ml/L at 0 C and 7.5 mg/L at 30 C, and in seawater the corresponding values are 11.2 and 6.1 mg/L.

HUMAN EXPOSURES AND DOSIMETRY

1.1.5

7

Food Contaminants

Chemical contaminants of almost every conceivable kind can be found in most types of human food. Food can acquire these contaminants at any of several stages in its production, harvesting, processing, packaging, transportation, storage, cooking, and serving. In addition, there are many naturally occurring toxicants in foods as well as compounds that can become toxicants upon conversion by chemical reactions with other constituents or additives or by thermal or microbiological conversion reactions during processing, storage, or handling. Each food product has its own natural history. Most foods are formed by selective metabolic processes of plants and animals. In forming tissue, these processes can act either to enrich or to discriminate against specific toxicants in the environment. For animal products, where the flesh of interest in foods was derived from the consumption of other life forms, there are likely to be several stages of biological discrimination and, therefore, large differences between contaminant concentrations in the ambient air and/or water and the concentrations within the animals.

1.2 HUMAN EXPOSURES AND DOSIMETRY People can be exposed to chemicals in the environment in numerous ways. The chemicals can be inhaled, ingested, or taken up by and through the skin. Effects of concern can take place at the initial epithelial barrier, that is, the respiratory tract, the gastrointestinal (GI) tract, or the skin, or can occur in other organ systems after penetration and translocation by diffusion or transport by blood, lymph, and so on. As illustrated in Fig. 1.2, exposure and dose factors are intermediate steps in a larger continuum ranging from release of chemicals into an environmental medium to an ultimate health effect. Exposure is a key step in this continuum and a complex one. The concept of total human exposure has developed in recent years as essential to the appreciation of the nature and extent of environmental health hazards associated with ubiquitous chemicals at low levels.

FIGURE 1.2

Environmental and biological modifiers of human exposure and health responses.

8


It provides a framework for considering and evaluating the contribution to the total insult from dermal uptake, ingestion of food and drinking water, and inhaled doses from potentially important microenvironments such as workplace, home, transportation, recreational sites, and so on. More thorough discussions of this key concept have been prepared by Sexton and Ryan (1988), Lioy (1990), and the National Research Council (NRC, 1991). Guidelines for Exposure Assessment have been formalized by the U.S. Environmental Protection Agency (U.S. EPA, 1992). 1.3 CHEMICAL EXPOSURES AND DOSE TO TARGET TISSUES Toxic chemicals in the environment that reach sensitive tissues in the human body can cause discomfort, loss of function, and changes in structure leading to disease. This section addresses the pathways and transport rates of chemicals from environmental media to critical tissue sites as well as retention times at those sites. It is designed to provide a conceptual framework as well as brief discussions of: (1) the mechanisms for—and some quantitative data on—uptake from the environment; (2) translocation within the body, retention at target sites, and the influence of the physicochemical properties of the chemicals on these factors; (3) the patterns and pathways for exposure of humans to chemicals in environmental media; (4) the effects of chemicals at the cellular and organ levels; and (5) the influence of age, sex, size, habits, health status, and so on. An agreed on terminology is critically important when discussing the relationships between toxic chemicals in the environment and human health. The terms used in this book are defined below: Exposure: Contact with external environmental media containing the chemical of interest. For fluid media in contact with the skin or respiratory tract, both concentration and contact time are critical. For ingested material, concentration and amount consumed are important. Deposition: Capture of the chemical at a body surface site on skin, respiratory tract, or GI tract. Clearance: Translocation from a deposition site to a storage site or depot within the body, or elimination from the body. Retention: Presence of residual material at a deposition site or along a clearance pathway. Dose: Amount of chemical deposited on or translocated to a site on or within the body where toxic effects take place. Target tissue: A site within the body where toxic effects lead to damage or disease. Depending on the toxic effects of concern, a target tissue can extend from whole organs down to specific cells to sub-cellular constituents. Exposure surrogates or indices: Indirect measures of exposure, such as: (1) concentra tions in environmental media at times or places other than those directly encountered; (2) concentrations of the chemical of interest, a metabolite of the chemical, or an enzy me induced by the chemical in circulating or excreted body fluids; and (3) elevations in body burden as measured by external probes. In summary, exposure represents contact between a concentration of an agent in air, water, food, or other material and the person or population of interest. The agent is the source

CONCENTRATION OF TOXIC CHEMICALS IN HUMAN MICROENVIRONMENTS

9

of an internal dose to a critical cell, organ, or tissue. The magnitude of the dose depends on a number of factors: (1) the volumes inhaled or ingested; (2) the fractions of the inhaled or ingested material transferred across epithelial membranes of the skin, the respiratory tract, and the GI tract; (3) the fractions transported via circulating fluids to target tissues; and (4) the fractional uptake by the target tissues. Each of these factors can have considerable intersubject variability. Sources of variability include activity level, age, sex, and health status as well as such inherent variabilities as race and size. With chronic or repetitive exposures, other factors affect the dose of interest. When the retention at, or effects on, the target tissues are cumulative and clearance or recovery is slow, the dose of interest can be represented by cumulative uptake. However, when the agent is rapidly eliminated, or when its effects are rapidly and completely reversible on removal from exposure, rate of delivery may be the dose parameter of primary interest.

1.4 CONCENTRATION OF TOXIC CHEMICALS IN HUMAN MICROENVIRONMENTS The technology for sampling air, water, and food is relatively well developed, as are the technologies for sample separation from copollutants, media, and interferences and for quantitative analyses of the components of interest. However, knowing when, where, how long, and at which rate and frequency to sample to collect data relevant to the exposures of interest is difficult, and requires knowledge of temporal and spatial variability of exposure concentrations. Unfortunately, we seldom have enough information of these kinds to guide our sample collections. Many of these factors are discussed in detail in the chapters that follow as they apply to the specific environmental toxicants being discussed. 1.4.1

Water and Foods

Concentrations of environmental chemicals in food and drinking water are extremely variable, and there are further variations in the amounts consumed because of the extreme variability in dietary preferences and food sources. The number of foods for which up-to-date concentration data for specific chemicals are available is extremely limited. Relevant human dietary exposure data are sometimes available in terms of market basket survey analyses. In this approach, foods for a mixed diet are purchased, cleaned, processed, and prepared as for consumption, and one set of specific chemical analyses is done for the composite mixture that is consumed. The concentrations of chemicals in potable piped water supplies depend greatly on the source of the water and its treatment history. Surface waters from protected watersheds generally have low concentrations of both dissolved minerals and environmental chemicals. Well waters usually have low concentrations of bacteria and environmental chemicals, but often have high mineral concentrations. Poor waste disposal practices may contribute to ground water contamination, especially in areas of high population density. Treated surface waters from lakes and rivers in densely populated and/or industrialized areas usually contain a wide variety of dissolved organics and trace metals, the concentrations of which vary greatly with season (because of variable surface runoff), with proximity to pollutant sources, with upstream usage, and with treatment efficacy. Uptake of environmental chemicals in bathing waters across intact skin is usually minimal in comparison to uptake via inhalation or ingestion. It depends on both the

10


concentration in the fluid surrounding the skin surface and the polarity of the chemical, with more polar chemicals having less ability to penetrate the intact skin. Uptake via skin can be significant for occupational exposures to concentrated liquids or solids. 1.4.2

Air

Although chemical uptake through ingestion and the skin surface is generally intermittent, inhalation provides a continuous means of exposure. The important variables affecting the uptake of inhaled chemicals are the depth and frequency of inhalation and the concentration and physicochemical properties of the chemicals in the air. Exposures to airborne chemicals vary widely among inhalation microenvironments, the categories of which include workplace, residence, outdoor ambient air, transportation, recreation, and public spaces. There are also wide variations in exposure within each category, depending on the number and strength of the sources of the airborne chemicals, the volume and mixing characteristics of the air within the defined microenvironment, the rate of air exchange with the outdoor air, and the rate of loss to surfaces within the microenvironment. 1.4.3

Workplace

Exposures to airborne chemicals at work are extremely variable in terms of composition and concentration, depending on the materials being handled, the process design and operation, the kinds and degree of engineering controls applied to minimize release to the air, work practices followed, and personal protection provided. Workplace air monitoring often involves breathing zone sampling, generally with passive samplers for gases and vapors or with personal battery-powered extraction samplers for both gases and particles; these operate over periods of 1–8 h. Analyses of the samples collected can provide accurate measures of individual exposures to specific air contaminants. Workplace air monitoring is also frequently done with fixed-site samplers or direct reading instruments. However, air concentrations at fixed sites may differ substantially from those in the breathing zones of individual workers. The fixed-site data may be relatable to the breathing zone when appropriate intercomparisons can be made, but otherwise they represent crude surrogates of exposure. The characteristics of equipment used for air sampling in industry are described in detail in Air Sampling Instruments (ACGIH, 2001). 1.4.4

Residential

Airborne chemicals in residential microenvironments are attributable to their presence in the air infiltrating from out-of-doors and to their release from indoor sources. The latter include unvented cooking stoves and space heaters, cigarettes, consumer products, and volatile emissions from wallboard, textiles, carpets, and so on. Personal exposures to chloroform, largely from indoor residential sources, are illustrated in Fig. 1.3, and the influence of smoking in the home on indoor exposures to respirable particulate matter is illustrated in Fig. 1.4. Indoor sources can release enough nitrogen dioxide (NO2), fine particle mass (FPM), and formaldehyde (HCHO) that indoor concentrations for these chemicals can be much higher than those in ambient outdoor air. Furthermore, their contributions to the total human exposure are usually even greater, since people usually spend much more time at home than in the outdoor ambient air.

CONCENTRATION OF TOXIC CHEMICALS IN HUMAN MICROENVIRONMENTS

11

FIGURE 1.3 Estimated frequency distributions of personal air exposures to chloroform: outdoor air concentrations, and exhaled breath values in Elizabeth-Bayonne, NJ area. Note: Air values are 12-h integrated samples. Breath value was taken following the daytime air sample (6:00 a.m. to 6:00 p.m.). Outdoor air samples were taken near participants’ homes. Source: Wallace et al. (1985).

1.4.5

Outdoor Ambient Air

For pollutants having national ambient air quality standards (NAAQS), there is an extensive network of fixed-site monitors, generally on rooftops. Although these devices generate large volumes of data, the concentrations at these sites may differ substantially from the concentrations that people breathe, especially for tailpipe pollutants such as carbon monoxide (CO), and reactive chemicals, such as ozone (O3) and sulfur dioxide (SO2). Data for other toxic pollutants in the outdoor ambient air are not generally collected on as routine a basis.

FIGURE 1.4 Respirable particle concentrations, six U.S. cities, November 1976 to April 1978. Source: National Academy of Science (1981).

12

1.4.6


Transportation

Many people spend from 1/2 to 3 h each day in autos or mass transport as they go to work, to school, or shopping. Inhalation exposures to CO in vehicles and garages can represent a significant fraction of total CO exposures. 1.4.7

Recreation and Public Spaces

Recreational exposure while exercising may be important to total daily exposure because the increased respiratory ventilation associated with exercise can produce much more than proportional increases in delivered dose and functional responses. Spectators and athletes in closed arenas can be exposed to high concentration of pollutants. For example, Spengler et al. (1978) documented high exposures to CO at ice rinks from exhaust discharges by the icescraping machinery.

1.5 INHALATION EXPOSURES AND RESPIRATORY TRACT EFFECTS 1.5.1

Deposition and Absorption

The surface and systemic uptake of chemicals from inhaled air depend on both their physical and chemical properties and on the anatomy and pattern of respiration within the respiratory airways. The basic structure of the respiratory tract is illustrated in Fig. 1.5. The following discussion outlines some of the primary factors affecting the deposition and retention of inhaled chemicals. More comprehensive discussions are available in recent reviews (ICRP, 1994; NCRP, 1997; U.S. EPA, 1996). Figure 1.6, from the 1994 ICRP Report, summarizes the morphometry, cytology, histology, function, and structure of the human respiratory tract, while Fig. 1.7 shows the compartmental model developed by ICRP (1994) to summarize particle transport from the deposition sites within the respiratory tract. Gases and vapors rapidly contact airway surfaces by molecular diffusion. Surface uptake is limited for compounds that are relatively insoluble in water, such as O3. For such chemicals, the greatest uptake can be in the lung periphery, where the residence time and surface areas are the greatest. For more water-soluble gases, dissolution and/or reaction with surface fluids on the airways facilitates removal from the airstream. Highly water-soluble vapors, such as SO2, are almost completely removed in the airways of the head, and very little of them penetrates into lung airways. For airborne particles, the most critical parameter affecting patterns and efficiencies of surface deposition is particle size. The mechanisms for particle deposition within respiratory airways are illustrated in Fig. 1.8. Almost all of the mass of airborne particulate matter is found in particles with diameters greater than 0.1 mm. Such particles have diffusional displacements many orders of magnitude smaller than those of gas molecules, and they are small in relation to the sizes of the airways in which they are suspended. Thus, the penetration of airborne particles into the lung airways is determined primarily by convective flow; that is, the motion of the air in which the particles are suspended. Some deposition by diffusion does occur for particles 0.5 mm, deposition by sedimentation occurs in small to midsized airways. For particles with aerodynamic diameters >2 mm, particle inertia is sufficient to cause particle motion to deviate from the flow streamlines, resulting in deposition by impaction on surfaces

INHALATION EXPOSURES AND RESPIRATORY TRACT EFFECTS

FIGURE 1.5

13

Structure of the respiratory tract. Reproduced from National Research Council (1979).

downstream of changes in flow direction, primarily in mid- to large-sized airways, which have the highest flow velocities. The concentration of deposition on limited surface areas within the large airways is of special interest with respect to dosimetry and the pathogenesis of chronic lung diseases such as bronchial cancer and bronchitis. Although particle inertia accounts for much of the “hot-spot” deposition on the trachea below the laryngeal jet and at the bifurcations of large lung airways, some of the concentrated deposition is attributable to inertial airflow, which directs a disproportionately large fraction of the flow volume toward such surfaces and, at the same time, lessens the boundary layer thickness. Thus, there is some preferential deposition of submicrometer-sized particles and gas molecules at small airway bifurcations. Quantitative aspects of particle deposition are summarized in Figs. 1.9–1.12. It can be seen that deposition efficiencies in the major structural–functional regions of the human respiratory tract are both strongly particle size dependent and highly variable among normal humans. Additional variability results from structural changes in the airways associated with disease processes. Generally, these involve airway narrowing or localized constrictions, which act to increase deposition and concentrate it on limited surface areas. All of the preceding was based on the assumption that each particle has a specific size. For particles that are hygroscopic, there is considerable growth in size as they take up water vapor

Interalveolar septa covered by squamous epithelium, containing capillaries, surfactant

Cuboidal alveolar epithelial cells (Type il. Surfactant-producing), covering 7% of alveolar surface area

†

**

**

16 – 18

Lymphatics

Alveolar sacs

Alveolar ducts

Respiratory bronchioles

Terminal bronchioles

Bronchioles

Bronchi

Trachea Main bronchi

Larynx

Nose Mouth

FIGURE 1.6

ET1

New

Esophagus

† LNTH

L

P

(T-B)

511

140m2

7.5 m2

4.5 x 107

4.6 x 105

2.6 x 10–1 m2 6.5 x 104

3 x 10–2 m2

—

—

Morphometry, cytology, histology, function, and structure of the respiratory tract and regions used in the 1994 ICRP dosimetry model.

Al

bb

BB

4.5 x 10–2 m2

2 x 10–3 m2

Zones Airway Number of Old* (Air) Location Surface Airways

Pharynx ET2 LNET (N-P) posterior

Anterior nasal passages

Anatomy

Lymph nodes are located only in BB region but drain the bronchial and alveolar interstitial regions as well as the bronchial region.

* Previous ICRP model. ** Unnumbered because of imprecise information.

Alveolar macrophages

Wall consists of alveolar entrance rings, squamous epithelial layer, surfactant

Squamous alveolar epithelium cells (Type i), covering 93% of alveolar surface areas

Gas exchange; very slow particle clearance

Mucous membrane, single-layer respiratory epithelium of cuboidal cells, smooth muscle layers

15

Mucous membrane, single-layer respiratory epithelium, less ciliated, smooth muscle layer

Respiratory epithelium consisting mainly of clara cells (secretory) and few ciliated cells

9 – 14

2–8

1

0

Mucous membrane, respiratory epithelium, no cartilage, no glands, smooth muscle layer

Mucous membrane, respiratory epithelium, cartilage plates, smooth muscle layer, glands

Mucous membrane, respiratory epithelium, cartilage rings, glands

Mucous membrane, respiratory or stratified epithelium, glands

Mucous membrane, respiratory epithelium (pseudostratified, ciliated, mucous), glands

Histology (Walls)

Air conduction; gas exchange; slow particle clearance

Respiratory epithelium with clara cells (No goblet cells) Cell types: - Ciliated cells - Nonciliated cells • Clara (secretory) cells

Air conditioning; Respiratory epithelium with goblet cells: temperature and Cell types: humidily, and - Ciliated cells cleaning; fast - Nonciliated cells: particle clearance; • Goblet cells air conduction • Mucous (secretory) cells • Serous cells • Brush cells • Endocrine cells • Basal cells • Intermediate cells

Cylology (Epithelium)

Regions used in Model

Extrathoracic Thoracic

Functions

Generation Number

Conditioning Conduction Gas-exchange transitory

0.175 x 10–3 m3 (Anatomical Dead Space) 0.2 x 10–3 m3 4.5 x 10–3 m3

Extrapulmonary Pulmonary

14


15

FIGURE 1.7 Compartment model to represent time-dependent particle transport from each region in 1994 ICRP model. Particle transport rate constants shown beside the arrows are reference values in d 1. Compartment numbers (shown in the lower right-hand corner of each compartment box) are used to define clearance pathways. Thus, the particle transport rate from bb1 to BB1 is denoted m4,7 and has the value 2 d 1.

in the airways. Some hygroscopic growth curves for acidic and ambient aerosols are illustrated in Fig. 1.13. Materials that dissolve into the mucus of the conductive airways or the surfactant layer of the alveolar region can rapidly diffuse into the underlying epithelia and the circulating blood, thereby gaining access to tissues throughout the body. Chemical reactions and metabolic processes may occur within the lung fluids and cells, limiting access of the inhaled material to the bloodstream and creating reaction products with either greater or lesser solubility and biological activity. Few generalizations about absorption rates are possible.

FIGURE 1.8 Schematic of mechanism for particle deposition in respiratory airways. Source: Lippmann and Schlesinger (1984).

16


1.0

0.8

ηn

Landahl & Tracewell Pattle Lippmann Hounam et al. Giacomelli-Maltoni et al. Martens & Jacobi Rudolf

1949 1961 1970 1971 1972 1973 1975

0.6 0.4 0.2

0 100

101

102

103

104

105

106

dae2 Q (µm2 cm3/s1)

FIGURE 1.9 Inspiratory deposition of the human nose as a function of particle aerodynamic 2 diameter and flow rate (dae Q). From: EPA (1997).

1.5.2

Translocation and Retention

Particles that do not dissolve at deposition sites can be translocated to remote retention sites by passive and active clearance processes. Passive transport depends on movement on or in surface fluids lining the airways. There is a continual proximal flow of lung surfactant from alveolar epithelial cells to and onto the mucociliary escalator, which begins at the terminal bronchioles, where it mixes with secretions from Clara and goblet cells in the airway epithelium. Within midsized and larger airways there are additional secretions from goblet cells and mucus glands, producing a thicker mucous layer having a serous subphase and an

FIGURE 1.10 Inspiratory extrathoracic deposition data in humans during mouth breathing as a 1=4 2 function of particle aerodynamic diameter, flow rate, and tidal volume (dae Q 2=3 VT ). From: EPA (1997).


17

FIGURE 1.11 Tracheobronchial deposition data in humans at mouth breathing as a function of particle aerodynamic diameter (dae). The solid curve represents the approximate mean of all the experimental data; the broken curve represents the mean excluding the data of Stahlhofen et al. From: EPA (1997).

overlying more viscous gel layer. The gel layer, lying above the tips of the synchronously beating cilia, is found in discrete plaques in smaller airways and becomes more of a continuous layer in the larger airways. The mucus reaching the larynx and the particles carried by it are swallowed and enter the GI tract. The total transit time for particles depositing on terminal bronchioles varies from 2 to 24 h in healthy humans, accounting for the relatively rapid bronchial clearance phase. Macrophage-mediated particle clearance via the bronchial tree takes place over a period of

FIGURE 1.12 Alveolar deposition data in humans as a function of particle aerodynamic diameter (dae). The solid curve represents the mean of all the data; the broken curve is an estimate of deposition for nose breathing by Lippmann (1977). From: EPA (1997).

18


FIGURE 1.13 Tracheobronchial particle deposition as a function of particle size at various ages for both stable iron oxide particles and hygroscopic sulfuric acid droplets that grow in size in the warm moist respiratory airways. Source: Martonen (1990).

several weeks. The particles depositing in alveolar zone airways are ingested by alveolar macrophages within about 6 h, but the movement of the particle-laden macrophages depends on the several weeks that it takes for the normal turnover of the resident macrophage population. At the end of several weeks, the particles not cleared to the bronchial tree via macrophages have been incorporated into epithelial and interstitial cells, from which they are slowly cleared by dissolution and/or as particles via lymphatic drainage pathways, passing through pleural and eventually hilar and tracheal lymph nodes. Clearance times for these later phases depend strongly on the chemical nature of the particles and their sizes, with half-times ranging from about 30 to 1000 days, or more. All of the characteristic clearance times cited refer to inert, nontoxic particles in healthy lungs. Toxicants can drastically alter clearance times. Inhaled materials affecting mucociliary clearance rates include cigarette smoke (Albert et al., 1974, 1975), sulfuric acid (H2SO4) (Lippmann et al., 1982; Schlesinger et al., 1983), O3 (Phalen et al., 1980; Schlesinger and Driscoll, 1987), SO2 (Wolff et al., 1977), and formaldehyde (Morgan et al., 1984). Macrophage-mediated alveolar clearance is affected by SO2 (Ferin and Leach, 1973), NO2 and H2SO4 (Schlesinger et al., 1988), O3 (Phalen et al., 1980; Schlesinger et al., 1988), and silica dust (Jammet et al., 1970). Cigarette smoke is known to affect the later phases of alveolar zone clearance in a dose-dependent manner (Bohning et al., 1982). Clearance pathways as well as rates can be altered by these toxicants, affecting the distribution of retained particles and their dosimetry.

1.6 INGESTION EXPOSURES AND GASTROINTESTINAL TRACT EFFECTS Chemical contaminants in drinking water or food reach human tissues via the GI tract. Ingestion may also contribute to uptake of chemicals that were initially inhaled, since material deposited on or dissolved in the bronchial mucous blanket is eventually swallowed. The GI tract may be considered a tube running through the body, the contents of which are actually external to the body. Unless the ingested material affects the tract itself, any systemic response depends on absorption through the mucosal cells lining the lumen. Although

SKIN EXPOSURE AND DERMAL EFFECTS

19

absorption may occur anywhere along the length of the GI tract, the main region for effective translocation is the small intestine. The enormous absorptive capacity of this organ results from the presence in the intestinal mucosa of projections, termed villi, each of which contains a network of capillaries; the villi result in a large effective total surface area for absorption. Although passive diffusion is the main absorptive process, active transport systems also allow essential lipid-insoluble nutrients and inorganic ions to cross the intestinal epithelium and are responsible for uptake of some contaminants. For example, lead may be absorbed via the system that normally transports calcium ions (Sobel et al., 1938). Small quantities of particulate material and certain large macromolecules, such as intact proteins, may be absorbed directly by the intestinal epithelium. Materials absorbed from the GI tract enter either the lymphatic system or the portal blood circulation; the latter carries material to the liver, from which it may be actively excreted into the bile or diffuse into the bile from the blood. The bile is subsequently secreted into the intestines. Thus, a cycle of translocation of a chemical from the intestine to the liver to bile and back to the intestines, known as the enterohepatic circulation, may be established. Enterohepatic circulation usually involves contaminants that undergo metabolic degradation in the liver. For example, DDT undergoes enterohepatc circulation; a product of its metabolism in the liver is excreted into the bile, at least in experimental animals (Hayes, 1965). Various factors serve to modify absorption from the GI tract, enhancing or depressing its barrier function. A decrease in gastrointestinal mobility generally favors increased absorption. Specific stomach contents and secretions may react with the contaminant, possibly changing it to a form with different physicochemical properties (e.g., solubility), or they may absorb it, altering the available chemical and changing translocation rates. The size of ingested particulates also affects absorption. Since the rate of dissolution is inversely proportional to particle size, large particles are absorbed to a lesser degree, especially if they are of a fairly insoluble material in the first place. For example, arsenic trioxide is more hazardous when ingested as a finely divided powder than as a coarse powder (Schwartz, 1923). Certain chemicals, for example, chelating agents such as EDTA, also cause a nonspecific increase in absorption of many materials. As a defense, spastic contractions in the stomach and intestine may serve to eliminate noxious agents via vomiting or by acceleration of the transit of feces through the GI tract.

1.7 SKIN EXPOSURE AND DERMAL EFFECTS The skin is generally an effective barrier against the entry of environmental chemicals. In order to be absorbed via this route (percutaneous absorption), an agent must traverse a number of cellular layers before gaining access to the general circulation (Fig. 1.14). The skin consists of two structural regions, the epidermis and the dermis, which rest on connective tissue. The epidermis consists of a number of layers of cells and has varying thickness depending on the region of the body; the outermost layer is composed of keratinized cells. The dermis contains blood vessels, hair follicles, sebaceous and sweat glands, and nerve endings. The epidermis represents the primary barrier to percutaneous absorption, the dermis being freely permeable to many materials. Passage through the epidermis occurs by passive diffusion. The main factors that affect percutaneous absorption are degree of lipid solubility of the chemicals, site on the body, local blood flow, and skin temperature. Some environmental chemicals that are readily absorbed through the skin are phenol, carbon tetrachloride,

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FIGURE 1.14

Idealized section of skin. Source: Birmingham (1973).

tetraethyl lead, and organophosphate pesticides. Certain chemicals, for example, dimethyl sulfoxide (DMSO) and formic acid, alter the integrity of skin and facilitate penetration of other materials by increasing the permeability of the stratum corneum. Moderate changes in permeability may also result following topical applications of acetone, methyl alcohol, and ethyl alcohol. In addition, cutaneous injury may enhance percutaneous absorption. Interspecies differences in percutaneous absorption are responsible for the selective toxicity of many insecticides. For example, chlorinated hydrocarbons (HC are about equally hazardous to insects and mammals if ingested but are much less hazardous to mammals when applied to the skin. This is because of their poor absorption through mammalian skin compared to their ready passage through the insect exoskeleton. Although the main route of percutaneous absorption is through the epidermal cells, some chemicals may follow an appendageal route, that is, entering through hair follicles, sweat glands, or sebaceous glands. Cuts and abrasions of the skin can provide additional pathways for penetration. 1.8 ABSORPTION THROUGH MEMBRANES AND SYSTEMIC CIRCULATION Depending upon its specific nature, a chemical contaminant may exert its toxic action at various sites in the body. At a portal of entry—the respiratory tract, GI tract, or skin—the chemical may have a topical effect. However, for actions at sites other than the portal, the agent must be absorbed through one or more body membranes and enter the general circulation, from which it may become available to affect cells and internal tissues (including the blood itself). The ultimate distribution of any chemical contaminant in the body is, therefore, highly dependent on its ability to traverse biological membranes. There are two main types of processes by which this occurs: passive transport and active transport. Passive transport is absorption according to purely physical processes, such as osmosis; the cell has no active role in transfer across the membrane. Since biological membranes contain lipids, they are highly permeable to lipid-soluble, nonpolar or

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nonionized agents and less so to lipid-insoluble, polar, or ionized materials. Many chemicals may exist in both lipid-soluble and -insoluble forms; the former is the prime determinant of the passive permeability properties for the specific agent. Active transport involves specialized mechanisms, with cells actively participating in transfer across membranes. These mechanisms include carrier systems within the membrane and active processes of cellular ingestion; that is, phagocytosis and pinocytosis. Phagocytosis is the ingestion of solid particles, whereas pinocytosis refers to the ingestion of fluid containing no visible solid material. Lipid-insoluble materials are often taken up by active-transport processes. Although some of these mechanisms are highly specific, if the chemical structure of a contaminant is similar to that of an endogeneous substrate, the former may be transported as well. In addition to its lipid-solubility characteristics, the distribution of a chemical contaminant is also dependent on its affinity for specific tissues or tissue components. Internal distribution may vary with time after exposure. For example, immediately following absorption into the blood, inorganic lead is found to localize in the liver, the kidney, and in red blood cells. Two hours later, about 50% is in the liver. A month later, approximately 90% of the remaining lead is localized in bone (Hammond, 1969). Once in the general circulation, a contaminent may be translocated throughout the body. In this process it may(1) become bound to macromolecules, (2) undergo metabolic transformation (biotransformation), (3) be deposited for storage in depots that may or may not be the sites of its toxic action, or (4) be excreted. Toxic effects may occur at any of several sites. The biological action of a contaminant may be terminated by storage, metabolic transformation, or excretion, the latter being the most permanent form of removal.

1.9 ACCUMULATION IN TARGET TISSUES AND DOSIMETRIC MODELS Some chemicals tend to concentrate in specific tissues because of physicochemial properties such as selective solubility or selective absorption on or combination with macromolecules such as proteins. Storage of a chemical often occurs when the rate of exposure is greater than the rate of metabolism and/or excretion. Storage or binding sites may not be the sites of toxic action. For example, CO produces its effects by binding with hemoglobin in red blood cells; on the contrary, inorganic Pb is stored primarily in bone but exerts it toxic effects mainly on the soft tissues of the body. If the storage site is not the site of toxic action, selective sequestration may be a protective mechanism, since only the freely circulating form of the contaminant produces harmful effects. Until the storage sites are saturated, a buildup of free chemical may be prevented. On the contrary, selective storage limits the amount of contaminant that is excreted. Since bound or stored toxicants are in equilibrium with their free form, as the contaminant is excreted or metabolized, it is released from the storage site. Contaminants that are stored (e.g., DDT) may remain in the body for years without effect. On the contrary, accumulation may produce illnesses that develop slowly, as occurs in chronic Cd poisoning. A number of descriptive and mathematical models have been developed to permit estimation from knowledge of exposure and one or more of the following factors: translocation, metabolism, and effects at the site of toxic action.

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The use of these models for airborne particulate matter generally requires knowledge of the concentration within specific particle size intervals or of the particle size distribution of the compounds of interest. Simple deposition models break the respiratory tract into regions (summarized by Vincent, 1999): Head airways, nasopharynx, extrathoracic: nose, mouth, nasopharynx, oropharynx, laryngopharynx. Tracheobronchial: larynx, trachea, bronchi, bronchioles (to terminal bronchioles). Gas exchange, pulmonary, alveolar: respiratory bronchioles, alveolar ducts, alveolar sacs, alveoli. Size-selective aerosol sampling can mimic the head airways and tracheobronchial airway regions so that airborne particle collection can be limited to the size fraction directly related to the potential for disease. More complex models requiring data on translocation and metabolism have been developed for inhaled and ingested radionuclides by the International Commission on Radiological Protection (ICRP, 1966, 1979, 1981, 1994).

1.10 INDIRECT MEASURES OF PAST EXPOSURES Documented effects of environmental chemicals on humans seldom contain quantitative exposure data and only occasionally include more than crude exposure rankings based on known contact with or proximity to the materials believed to have caused the effects. Reasonable interpretation of the available human experience requires some appreciation of the uses and limitations of the data used to estimate the exposure side of the exposure– response relationship. The discussion that follows is an attempt to provide background for interpreting data, and for specifying the kinds of data needed for various analyses. Both direct and indirect exposure data can be used to rank exposed individuals by exposure intensity. External exposure can be measured directly by collection and analysis of environmental media. Internal exposure can be estimated from analyses of biological fluids and in vivo retention. Indirect measures generally rely on work or residential histories with some knowledge of exposure intensity at each exposure site and/or some enumeration of the frequency of process upsets and/or effluent discharges that result in high-intensity short-term exposures. 1.10.1

Concentrations in Air, Water, and Food

Historic data may occasionally be available on the concentrations of materials of interest in environmental media. However, they may or may not relate to the exposures of interest. Among the more important questions to be addressed in attempts to use much data are: (1) How accurate and reliable were the sampling and analytical techniques used in the collection of the data? Were they subjected to any quality assurance protocols? Were standardized and/or reliable techniques used? (2) When and where were the samples collected, and how did they relate to exposures at other sites? Air concentrations measured at fixed (area) sites in industry may be

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much lower than those occurring in the breathing zone of workers close to the contaminant sources. Air concentrations at fixed (generally elevated) community air-sampling sites can be either much higher or much lower than those at street level and indoors as a result of strong gradients in source and sink strengths in indoor and outdoor air. (3) What is known or assumed about the ingestion of food and/or water containing the measured concentrations of the contaminants of interest? Time at home and dietary patterns are highly variable among populations at risk. 1.10.2

Biological Sampling Data

Many of the same questions that apply to the interpretation of environmental media concentration data also apply to biological samples, especially quality assurance. The time of sampling is especially critical in relation to the times of the exposures and to the metabolic rates and pathways. In most cases, it is quite difficult to separate the contributions to the concentrations in circulating fluids of levels from recent exposures and those from long-term reservoirs. 1.10.3

Exposure Histories

Exposure histories per se are generally unavailable, except in the sense that work histories or residential histories can be interpreted in terms of exposure histories. Job histories, as discussed below, are often available in company and/or union records and can be converted into relative rankings of exposure groups with the aid of long-term employees and managers familiar with the work processes, history of process changes, material handled, tasks performed, and the engineering controls of exposure. Routine, steady-state exposures may be the most important and dominant exposures of interest in many cases. On the contrary, for some health effects, the occasional or intermittent peak exposures may be of primary importance. In assessing or accumulating exposure histories or estimates, it is important to collect evidence for the frequency and magnitude of the occasional or intermittent releases associated with process upsets. 1.11 CHARACTERIZATION OF HEALTH 1.11.1

Definitions of Health

There is no universally accepted definition of health. Perhaps the most widely accepted one today is that of the World Health Organization, which describes health as a state of complete physical, mental, and social well-being, and not merely the absence of disease or infirmity. Unfortunately, by a strict interpretation of this rather idealistic definition, very few people could be considered healthy. The discussion to follow is limited largely to physical well-being. The health effects discussed are those that can be recognized by clinical signs, symptoms, or decrements in functional performance. Thus, for all practical purposes, in this volume we consider health to be the absence of measurable disease, disability, or dysfunction. 1.11.2

Health Effects

Recognizable health effects in populations are generally divided into two categories: mortality and morbidity. The former refers to the number of deaths per unit of population

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per unit time, and to the ages at death. Morbidity refers to nonfatal cases of reportable disease. Accidents, infectious diseases, and massive overexposures to toxic chemicals can cause excess deaths to occur within a short time after the exposure to the hazard. They can also result in residual disease and/or dysfunction. In many cases, the causal relationships are well defined, and it may be possible to develop quantitative relationships between dose and subsequent response. The number of people exposed to chemical contaminants at low levels is, of course, much greater than the number exposed at levels high enough to produce overt responses. Furthermore, low-level exposures are often continuous or repetitive over periods of many years. The responses, if any, are likely to be nonspecific, for example, an increase in the frequency of chronic diseases that are also present in nonexposed populations. For example, any small increase in the incidence of heart disease or lung cancer attributable to a specific chemical exposure would be difficult to detect, since these diseases are present at high levels in nonexposed populations. In smokers they are likely to be influenced more by cigarette exposure than by the chemical in question. Increases in the incidence of diseases from low-level long-term exposure to environmental chemicals invariably occur among a very small percentage of the population and can only be determined by large-scale epidemiological studies (epidemiology is the study of the distribution and frequency of diseases in a specific population) involving thousands of person-years of exposure. The only exceptions are chemicals that produce very rare disease conditions, where the clustering of a relatively few cases may be sufficient to identify the causative agent. Notable examples of such special conditions are the industrial cases of chronic berylliosis caused by the inhalation of beryllium-containing dusts, a rare type of liver cancer that resulted from the inhalation of vinyl chloride vapors, and pleural cancers that resulted from the inhalation of asbestos fibers. If these exposures had produced more commonly seen diseases, the specific materials might never have been implicated as causative agents. Low-level chemical exposures may play contributory, rather than primary, roles in the causation of an increased disease incidence, or they may not express their effects without the co-action of other factors. For example, the excess incidence of lung cancer is very high in uranium miners and asbestos workers who smoke cigarettes but is only marginally elevated among nonsmoking workers with similar occupational exposures. For epidemiological studies to provide useful data, they must take appropriate account of smoking histories, age, and sex distributions, socioeconomic levels, and other factors that affect mortality rates and disease incidence. 1.11.2.1 Mortality In industrialized societies, there is generally good reporting of mortality and age at death but, with few exceptions, quite poor reporting of cause of death. In studies that are designed to determine associations between exposures and mortality rates, it is usually necessary to devote a major part of the effort to follow-up investigations of cause-of-death. The productivity of these follow-ups is often marginal, limiting the reliability of the overall study. 1.11.2.2 Morbidity Difficult as it may be to conduct good mortality studies, it is far more difficult, in most cases, to conduct studies involving other health effects. Although there is generally little significant variability in the definition of death, there is a great deal of variation in the diagnosis and reporting of many chronic diseases. There are variations

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between and within countries and states, and these are exacerbated by the differences in background and outlooks of the physicians making the individual diagnoses. Furthermore, there are some important chronic diseases that cannot be definitively diagnosed in vivo. Many epidemiological studies rely on standardized health status questionnaires, and the success of these studies depends heavily on the design of the questionnaires. Of equal importance in many studies are the training and motivation of the persons administering the questionnaires. Similar considerations apply to the measurement of functional impairment. The selection of the measurements to be used is very important; those functions measured should be capable of providing an index of the severity of the disease. Equally important here are the skills of the technicians administering these tests and their maintenance and periodic recalibration of the equipment. Some studies try to avoid bias from the administrators of the questionnaires and functional tests by having the selected population enter the desired information themselves. They may be asked to make appropriate notations in notebook diaries or to call a central station whenever they develop the symptoms of interest. Other investigations use nonsubjective indices such as hospital admissions, clinic visits, and industrial absenteeism as their indicators of the health effects to be associated with the environmental variables.

1.12 EXPOSURE–RESPONSE RELATIONSHIPS Exposure–response relationships can be developed from human experience, but there are many chemicals that are known to be toxic in animals for which the extent of human toxicity, if any, is unknown. In order to use animal bioassay data for the prediction of human responses to environmental exposures, it is necessary to make two major kinds of extrapolation. One is determine or estimate the relative responsiveness of humans and the animal species used in the bioassays. The second is to extrapolate from the observed effects resulting from relatively high administered doses to the much lower levels of effects still of concern at much lower levels of environmental exposure. To deal with interspecies extrapolation, estimates are made on the basis of whatever is known about differences in uptake from environmental media, metabolic rates and pathways, retention times in target tissues, and so on, and tissue sensitivities. As uncertain as these extrapolations are, they are more straightforward than the low-dose extrapolation. The goal of the dose–response assessment is to predict what response, if any, might occur, 10- to 1000-fold below the lowest dose tested in rodents (this is more representative of the range of doses to which humans are usually exposed). Because it would require the testing of thousands of animals to observe a response at low doses, mathematical models are used (Munro and Krewski, 1981). To appreciate the level of uncertainty in the dose extrapolation process and the typical regulatory use of low-dose models, it is useful to discuss the dose– response curve. However, reliance on the results of only one mathematical model is a potential pitfall in the dose–response assessment. There are at least six different modeling approaches that may need to be considered when estimating the risks at low doses. These models include the probit, multihit, multistage, Weibull, one-hit, and the Moolgavkar–Knudson–Venzon (MKV) biologically based approaches (Moolgavkar et al., 1988). Nearly all of them can yield results that are plausible. No single statistical model can be expected to predict accurately the low-dose response with greater certainty than another. As discussed by Paustenbach (1990),

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FIGURE 1.15 The fit of most dose–response models to data in the range tested in animal studies is generally similar. However, because of the differences in the assumptions on which the equations are based, the risk estimates at low doses can vary dramatically between different models. Source: Paustenbach (1990).

one possible way to resolve this problem is to present the best estimate of the risk from the two or three models that are considered equally reasonable along with the upper- and lower-bound estimates. An alternate approach is to identify a single value based on the “weight of evidence,” as the EPA did for dioxin (U.S. EPA, 1988) Low-dose models usually fit the rodent data in the dose region used in the animal tests. However, they often predict quite different results in the unobserved low-dose region (Fig. 1.15). The results of the most commonly used low-dose models usually vary in a predictable manner because the models are based on different mathematical equations for describing the chemical’s likely behavior in the low-dose region. In general, the scientific underpinnings of the dose–response models are based on the present understanding of the cancer process caused by exposure to ionizing radiation and genotoxic chemicals (NRC, 1980). Both types of agents may well have a linear, or a nearly linear, response in the low-dose region. However, promoters and cytotoxicants (e.g., nongenotoxicants) would be expected to be very nonlinear at low doses and may have a genuine or practical threshold (a dose below which no response would be present) (Squire, 1987; Butterworth and Slaga, 1987). Thus, the linearized multistage model may be inappropriate for dioxin, thyroid-type carcinogens, nitrolotriacetic acid, and, presumably, similar nongenotoxic chemicals (Paynter et al., 1988; Andersen and Alden, 1989). For these types of chemicals, the MKV model, or one of the other biologically based models, should be more appropriate (Moolgavkar, 1978; Ellwein and Cohen, 1988). 1.12.1

Summary of Exposure- and Dose-Related Responses

Studies of the specific responses of biological systems to varying levels of exposure can provide a great deal of information on the nature of the responses, their underlying


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FIGURE 1.16 Dose (population) response relationship with suggested distinction between basic (toxicological) and practical (health) scales on the three axes. The illustrative curve on the horizontal plane portrays the dose–response relationship for the middle (50%) of the exposed population; the curve on the vertical plane shows the percentages of population response of the indicated degree over the whole range of doses. The vertical line from the dose scale indicates the magnitude of dose needed to produce the indicated degree of response at the 50% population level. Source: Hatch (1968).

causes, and the possible consequences of various levels of exposure. However, it must be remembered that the data are most reliable only for the conditions of the test and for the levels of exposure that produced clear-cut responses. Generally, in applying experimental data to low-level environmental exposure conditions, it is necessary to extrapolate to delivered doses that are orders of magnitude smaller than those that produced the effects in the test system. Since the slope of the curve becomes increasingly uncertain the further one extends it beyond the range of experimental data, the extrapolated effects estimate may be in error by a very large factor. The basic dimensions of the dose–response relationship for populations were described by Hatch (1968), as illustrated in Fig. 1.16. Many factors affect each of the basic dimensions. 1.12.1.1 Factors Affecting Dose The effective dose is the amount of toxicant reaching a critical site in the body. It is proportional to the concentrations available in the environment: in the air breathed, the water and food ingested, and so on. However, the uptake also depends on the route of entry into the body and the physical and chemical forms of the contaminant. For airborne contaminants, for example, the dose to the respiratory tract depends on whether they are present in a gaseous form or as an aerosol. For contaminants that are ingested, uptake depends on transport through the membranes lining the gastrointestinal tract and, in turn, is dependent on both aqueous and lipid solubilities.

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For contaminants that penetrate membranes, reach the blood, and are transported systemically, subsequent retention in the body depends on their metabolism and toxicity in the various tissues in which they are deposited. In all of these factors, there are great variations within and between species, and therefore great variations in effective dose for a given environmental level of contamination. 1.12.1.2 Factors Affecting Response The response of an organism to a given environmental exposure can also be quite variable. It can be influenced by age, sex, the level of activity at the time of exposure, metabolism, and the competence of the various defense mechanisms of the body. The competence of the body’s defenses may, in turn, be influenced by the prior history of exposures to chemicals having similar effects, since those exposures may have reduced the reserve capacity of some important functions. The response may also depend on other environmental factors, such as heat stress and nutritional deficiencies. These must all be kept in mind in interpreting the outcomes of controlled exposures and epidemiological data and in extrapolating results to different species and across various age ranges, states of health, and so on. 1.12.1.3 Factors Affecting Individual Susceptibility The complete evaluation of the pathogenesis of human disease requires identification and assessment of the genetic, lifestyle, and environmental risk factors. Environmental factors are clearly critical to disease prevention because, at the societal level, they are determined by the most controllable processes. Thus, the overall purpose of environmental medicine is to improve our knowledge of the more commonly encountered environmental agents that present the greatest concern to human health. In the past, diseases were categorized by the main etiological factor into either (a) genetic, (b) lifestyle-induced, or (c) environmental-induced disorders. Illustrative examples for each category include (a) alpha-1 antitrypsin [serine (or cysteine) proteinase inhibitor, clade A, member1] deficiency-induced chronic obstructive pulmonary disease; (b) cigarette-induced lung cancer; and (c) asbestos-induced mesothelioma, respectively. Environmental medicine almost exclusively focused on the latter category, and has had its best impacts when armed with knowledge of quantifiable exposure (e.g., occupational diseases). Fortunately, additional cases of occupational diseases (e.g., mining-related coal workers pneumoconiosis or asbestosis) are becoming increasingly rare among the general population due to improved work practices. However, while categorizations based on etiological factors have allowed the development of therapeutic strategies to treat disease, they also may have limited past attempts to prevent disease. Disease prevention has a greater ability than most common treatment modalities to reduce incidence and to extend life expectancy. In the illustrative examples above, preventative approaches would include: (a) genetic counseling and possibly genereplacement therapy; (b) public health education and smoking cessation pharmaceuticals; and (c) reduction or elimination (banning widespread usage) of asbestos exposure. Each of theses approaches have had tremendous value in focusing efforts on disease causalities and thus has had dramatic impacts on disease prevention. Although the application of these simple approaches to disease prevention has been partially successful, further success will require more sophisticated approaches based on a deeper understanding of disease pathogenesis. With the widespread use of genetic screening, it soon became apparent that homozygotic recessive carriers (individuals inheriting both copies of the disease susceptibility alleles) often did not developed major signs and


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symptoms of the “genetic” disease. For example, homozygotic twins do not have 100% concordance in disease outcomes. Similarly, disease discordance is noted in lifestyleinduced diseases. For example, among individuals with exposures exceeding 50 pack-years, only four out five cigarette smokers develop a tobacco-related disease and these diseases vary in the population (i.e., multisite cancer, cardiovascular, and respiratory disease). Likewise, although diseases such as asbestosis, lung cancer, and mesthelioma are enriched in populations with excessive asbestos exposure, the incidence of affected workers given equivalent exposures in not 100%. What can explain a lack of penetrance (affected individuals/genotype positive individuals) in equally susceptible, equally risky, equally exposed populations? It is the interactions of additional factors within and among each etiological category. Common pathological conditions are complex and involve multiple rather than a single gene(s). Single gene diseases with strongly expressivity typically appear early in life, but due to a lack of complete penetrance affect only a few members of the population. Alternatively, most diseases involve multiple gene–gene interactions and are present in a large percentage of the population. Genetic polymorphisms that are strongly fixed in the genome probably did not arise from modern lifestyle or environmental factors (e.g., coal mining), but from ancient lifestyle factors or infectious agents. Another reason why most common diseases are complex is because the selective advantage to a population that stabilizes a genetic polymorphism is likely to be dependent on multiple alterations in multiple genes. Many polymorphisms also may have been acquired from phylogenetic ancestry (which are likely to be greater than those that uniquely arise within a given species). Thus, the mechanisms by which bacteria combat virus, or how drosophila resist bacteria (toll-receptors), or how mice resist influenza, may be conserved throughout species. This situation is further complicated by the fact that genes that impart sensitivity may also impart resistance to another disease. For example, the protection from tuberculosis may be advantageous to the population, but may impart increased susceptibility to chronic inflammatory diseases, like arthritis, to a portion of the population. From a solely genetic stand point, it would be advantageous if multiple genes contribute to a survival phenotype from a severe disease entity [i.e., the fully developed phenotype being dependent on gene–gene interactions (epistasis)]. Phenotypes with complex gene interaction require only a few members of an outbred population (humans) to carry the exact set of all the resistant alleles. Phenotypes dependent upon multiple genes thereby reduce the negative consequences of the combinatorial effects of multiple alleles to only a few individuals. Several other members of a population could inherit in partial combinations that would have little observable phenotypic expression. Thus, while only a few members of the population might survive an infectious epidemic, many members of population share the risky alleles. With the initial assembling of the entire human genome by the Human Genome Project, substantial research has been invested in the identification of all disease causing genetic polymorphisms. However, individual susceptibility to common diseases is not solely controlled by multiple gene–gene interactions. Rather disease penetrance is also influenced by multiple gene–lifestyle, gene–environment, lifestyle–environmental, and gene– lifestyle–environmental factors. Using our illustrative disease example, clear synergistic interactions occur among individuals who have: (a) alpha-1 antitrypsin deficiency and smoke cigarettes (gene–lifestyle interaction); (b) glutathione S-transferase pi 1 deficiency and are exposed to environmental tobacco smoke or to excessive air pollution (ozone and particulate matter) (gene-environment interaction); or (c) smoke and work with asbestos or radon (lifestyle–environment interactions).

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1.12.2 Genomic Approaches to Understanding Gene-Lifestyle-Environmental Factors in Complex Disease Pathogenesis Environmental health sciences have developed a wonderful cadre of tools to obtain global (nearly complete) evaluations of the genetic variants, transcriptional profile, protein usage and activation state, and the metabolic capability of individuals and populations following environmental stress. Using high-throughput microarray or fluidic luminesence systems, genomics seeks to evaluate the entire genetic makeup of an individual and thereby identify candidate gene suspected to have a role in determining susceptibility. Genomics is based on known biological functional, cellular location, or pathophysiological roles of a gene, and seeks to identify the allelic variants that associate with increased risk. Moreover, the mapping of over 150,000 single nucleotide polymorphisms (SNPs) throughout the genome allows nearly complete coverage of all human variability and through linkage disequilibrium identification of small chromosomal region (areas of a few tens of thousands base pairs). Particularly attention is focused on nonsynonymous SNPs that result in alteration of RNA recognition codons and lead to amino acid changes in the predicted protein. Because regions of chromosomes are often inherited together, the number of independently inherited SNPs is reduced. Thus, the reduction to informative differences (tag-SNPs) make these analyses even more powerful and essentially entire genome coverage is possible. Supportive of genome wide linkage analysis is transcriptomics, proteomics, and metabonomics. Transcriptomics studies large sets of messenger ribonucleic acid (mRNA) molecules, or transcripts produced by cells in culture (revealing cell-type specificity), isolated tissue, or a whole organism. The transcriptome, unlike the genome, which is fixed (excluding acquired mutations), can be altered by the environment and reflects the cell’s attempt to acclimate to adverse conditions. The supportive high-throughput technology includes DNA microarrays that typically monitor steady-state transcript levels of 30,000 genes. This includes essentially all known genes and genes yet to be fully understood and annotated (e.g., predicted gene products and expressed sequence tags). Confirmational approaches to key genes identified by microarray include quantitative reverse transcription polymerase chain reaction, ribonuclease protection assay, and Northern blot that are conducted on single or small sets of transcripts. Like the transcriptome, the proteomics is a global approach to evaluate altered protein usage and activation state induced by environmental signaling. Indeed, it includes many signaling peptides (e.g., kinases) that generate amplifying cascades that alter cell functions including motility, transcription, cell–cell communication, proliferation, immunity, and apoptosis. The proteome includes nascent propeptides, mature inactive peptides, activated peptides, and peptide marked secretion or degradation. While the number of genes and possible transcripts are estimated to be less than 35,000 in humans, the proteome may have over 1 million members. Moreover, proteomics focuses on protein–protein, and protein–macromolecular interactions, and is thus closer to functional significance than genomics or transcriptomics. High-throughput technologies supporting proteomics include, gas (gas chromatography), fluidic (high-pressure liquid chromatography) or gel (electrophoresis) separation and large scale, mass spectrometric protein identification. Confirmational approaches include immuno (Western) blot, antibody arrays, and enzyme-linked immunosorbent assays (ELISA), and radioimmunoassay. Also under rapid development, metabonomics (or metabolomics) is the global approach to the assessment of the metabolic response of living systems to environmental stimuli. Typically the variety of small molecule metabolites generated by a living system is relatively

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small (80% predicted, FEV1 25 mg/m3). This exposure range, however, is outside the scope of this review. From the field investigations, it appears that although the symptoms observed are not systematically described, they are more frequent among those exposed than among the

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nonexposed. It was found that complaints seem to be present when the concentrations exceed 1.7 mg/m3. Below 1.7 mg/m3 complaints may occur if other types of simultaneous exposures are present (Mølhave, 1986). The concentrations reported from field investigations are improperly documented, and they may be biased. The exposure range of 0.19–0.66 mg/m3 was estimated in the Danish Town Hall Study for the lower, no-effect threshold (Zweers et al., 1990). This range corresponds to the range of lower limit of concentrations in buildings with complaints (Mølhave, 1986) and is at present the best first estimate of the lower exposure limit for no effects of VOC, but should, of this level of documentation, only be used for screening purposes. The laboratory experiments indicate that the main effects can be experimentally reproduced and acutely follow the exposure. No field investigations have been reported of tests of the effects of elimination or modification of VOC exposure. Postexposure measurements during the controlled experiments indicate, however, that the effects are reversible and disappear shortly after exposure (Mølhave et al., 1991). Exposures in most field investigations are multifactorial, as factors other than VOC exposure may exceed their no-effects levels and most of the effects reported in field investigations may have more than one cause. It is, therefore, not surprising that effect of VOC exposures in field investigations seem to occur at lower exposure levels than in controlled experiments, where most other factors are supposed to be below their noeffect levels. In the experiments, the exposure times were less than 3 h, which, from field experience, seem too short to cause severe subacute effects at low exposure levels. More research is needed to test if subacute effects may occur after prolonged exposures. In conclusion, there is no evidence to contradict the proposed causal link between low-level exposure to VOC and the effects shown in Table 8.8. On the contrary, evidence from both field investigations and controlled exposure experiments supports causality. The field investigations and controlled experiments as concluded by the Nordic consensus group (Andersson et al., 1997) are, however, still too few to allow a final conclusion.

8.6 GUIDELINES FOR VOLATILE ORGANIC COMPOUNDS IN NONINDUSTRIAL INDOOR ENVIRONMENTS-PRINCIPLES FOR ESTABLISHMENT OF GUIDELINES Two different concepts seem to be used by different authors in the evaluation of indoor air. The first refers to the quantitative evaluation of the risks of adverse irreversible health effects (e.g., asphyxiation by CO, or lung cancer associated with radon) or the risks of reversible or irreversible changes in the body’s physiological functions (e.g., nervous system effects). This is the traditional occupational or environmental evaluation of health risks, which is done according to standard toxicological principles including risk estimates and cost-benefit analysis. This is the background for air quality guidelines and ambient air standards for air pollutants and for TLVs and occupational exposure standards for light levels and sound levels, and is based on lists of high-risk compounds. The second concept refers to qualitative evaluations of the atmospheric environment and is, in many respects, a new concept for regulation that is often neglected in discussions of indoor air quality: Air deodorants, painting, wallpapers, or music can be liked by some persons and disliked by others, and generally accepted principles for regulation of the quality

252

VOLATILE ORGANIC COMPOUNDS AND SICK BUILDING SYNDROME

of the indoor environment with respect to odors, sounds, and colors may be impossible to establish, if indeed such regulation is wanted at all. Some general conclusions may, however, be drawn about the principles to apply in setting future regulations of VOCs in the indoor climate. Such principles can be extracted from existing regulations, for example, in building codes, for acoustics and lightning that contain additional qualitative concerns besides those used for the setting of TLV levels. For these guidelines the first basic principle implicit is that the building must support a specified range of human activities, habits, and preferences. Complaints and decreased performance will automatically follow if the occupants try to do activities outside this range, for example, reading in too dark a room with disturbing intermittent noise peaks or working with a video screen with many light reflections. This range of activity may be different for homes and offices, as the activity patterns in homes include recreation, rest, and sleep and other activities normally not found in offices. Further, the occupants of homes may be more sensitive than the working population, as they include the sick, young, and old fractions of the population. The second basic principle originates from the assumption that humans do not feel well if they do not have the optimal use of their senses to perceive their environment and to monitor the activities they are performing. The ideal indoor environment, therefore, seems to allow the occupants to use their senses to perceive their environment and to monitor the activities they are performing. The occupants should be able to use their senses to pick up wanted environmental signals undisturbed by exposures to unwanted information noise. This means that unwanted environmental information should be damped, such as the sound of a typing machine in an office or the neighbor’s radio in a home. On the contrary, our own conversation or the perception of sounds related to our own activities should be eased. In short, this second principle tells us to optimize the signal-to-noise ratio for our senses by allowing the wanted signals to propagate to the occupants and damping unwanted sensory signals. This principle, if true, explains why occupants have such different optimal environments. The signals that bear information to one person about his own activity and environment create sensory noise for the other person. Therefore, the signals relevant for one person differ from those relevant for his or her neighbor. To summarize the discussion on sensory perception of VOCs as air pollutants, the acceptable exposure range may be defined as follows. In normal rooms the air quality is acceptable if no unacceptable health risk exists, and if all sources for chemical stimulation can be identified by the occupants, and those sources bearing unwanted information can be removed, should the occupants desire to do so. 8.6.1

A Tentative Guideline for VOCs in Nonindustrial Environments

The observations summarized here are based on investigations that have major limitations. Presently, they do not form an acceptable basis for setting official recommendations or guidelines (Andersson et al., 1997), but discussions on the possible principles of such guidelines have been initiated (Mølhave, 1998). The observations do, however, indicate that VOCs may be important for indoor air quality, especially in the form of discomfort from odors, irritative symptoms in eyes, nose, and throat, and headache. The effects may also include effects related to productivity and performance. Such effects have not yet been positively identified.

GUIDELINES FOR VOLATILE ORGANIC COMPOUNDS

253

Most indoor environments in nonindustrial buildings are polluted with 10–100 VOCs as air pollutants, each generally in concentrations ranging from nanograms to few hundred micrograms per cubic meter. The list of compounds often identified indoors may include 1000 compounds. The toxicity of these compounds in the low exposure range is mostly unknown, and their possible interactions cannot be predicted with the existing knowledge. Lacking guidelines for indoor air quality, practitioners have had to use simple approximative tools for making evaluations of air quality. One of these is the TVOC, which is the sum of the individual concentrations of all VOC present. TVOC has been discussed by several international working groups (EU-ECA, 1997; Andersson et al., 1997) as described above. TVOC should only be used as a screening tool to identify potential unacceptable exposures to VOC, and should not be used as the only tool to make definitive conclusions. Tentative screening values for VOCs in nonindustrial environments have been offered (Mølhave, 1991). The tentative conclusion of the available epidemiological studies and the exposure experiments is that if the presence of specific potential irritants can be excluded (which they seldom can), irritation is unlikely to follow from exposure to VOCs below about 0.2 mg/m3. The ambient air levels are generally below that level, which for screening purposes may be used as a first estimate of a lower limit for possible effects of VOCs. At concentrations higher than about 3 mg/m3, complaints seemed to occur in most investigated buildings with occupants having symptoms. In controlled exposure experiments using on mixture of 22 VOCs, odors were significant at 3 mg/m3. At 5 mg/m3 in the same exposure experiments, objective effects were indicated in addition to subjective irritation. Exposures for 50 min to 8 mg/m3 led to significant irritation of mucous membranes in eyes, nose, and throat. Few indications are given in the literature that allows an estimate of the threshold for headache. Concentrations below 3 mg/m3 were found, in field investigation, to produce a significant difference in frequencies of headache between problem buildings and control buildings. On the contrary, significant headache was found in only one of the exposure experiments and then at 25 mg/m3. The reason for the lower threshold in field investigations may be either the interaction of other exposures, or the effect of longer exposure durations. Therefore, based on the present information, the threshold for TABLE 8.9 Tentative Dose–Response Relationship for Discomfort Resulting from Exposure to Solvent-like Volatile Organic Total Concentration (mg/m3) 25

Irritation and Discomfort No irritation or discomfort Irritation and discomfort possible if other exposures interact Exposure effect and probable headache possible if other exposures interact Additional neurotoxic effects other than headache may occur

Exposure Range The comfort range The multifactorial exposure range The discomfort range

The toxic range

The table only refers to VOCs with a normal range of biological activity and should only be used for screening purposes. Source: Mølhave (1991).

254


headache and other weak neurotoxic effects caused by exposure of less than a few hours duration are expected to be between 3 and 25 mg/m3. These conclusions refer to the most prevalent effects of VOCs on normal subjects. Risk groups may exist that will respond more strongly than the normal population. Furthermore, future investigations dealing with larger groups of persons may reveal special effects such as allergy or carcinogenicity associated with low-level exposures to VOCs. These special effects, however, have not been demonstrated for the type and concentration of VOCs found in indoor air. A tentative dose–response relationship for discomfort resulting from exposure to VOCs is shown in Table 8.9. The table, which should only be used for screening purposes, indicates a possible no-effect level at about 0.2 mg/m3. A multifactorial exposure range may exist from 0.2 to 3 mg/m3 in which odor irritation and discomfort may appear as consequences of VOC exposure depending on the types of compounds present and if other exposures contribute to the etiology. Above about 3 mg/m3, effects of VOC exposure are likely and exposures above 25 mg/m3 may be expected to cause toxic effects.

REFERENCES American Conference of Governmental Industrial Hygienists (ACGIH) (2006) TLVs: Threshold Limit Values and Biological Exposure Indices for 2006. Cincinnatti, USA:ACGIH. Andersson K, Bakke JV, Bj€orseth O, Bornhehag CG, Clausen G, Hongslo JK, Kjellman M, Kjœrgaard SS, Levy F, Mølhave L, Skerfving S, Sundell J (1997) TVOC and health in non-industrial indoor environments. Reports from a Nordic Scientific consensus meeting at La˚ngholmen in Stockholm, 1996. Indoor Air 7:78–91. Andrew LS, Snyder R (1980)Toxic effects of solvents and vapors. In:Casarett LJ, Doull J,editors. Casarett and Doull’s Toxicology. The Basic Science of Poisons. New York, USA: Macmillan. Ashford N, Hewinzow B, L€utjen K, Maruoli C, Mølhave L, M€onch B, Papadopoulos S, Rest K, Rosdahl D, Siskos P, Volonakis E (1994) Chemical sensitivity in selected European countries: an explorative study. Commission of the European Union, DG-V, agreement SOC 93 2027 48 05E01 (93CVVF1-613-0). Report from Ergonomia Ltd, Athens, Greece. Berglund B, Johansson I, Lindvall T (1981) Underlag f€or Ventilationsnormer. ETAPP II (Ventilation Requirements, in Swedish). Stockholm, Sweden: The National Institute of Environmental Medicine. Berglund B, Johansson I, Lindvall T (1982) A longitudinal study of air contaminants in a newly built preschool. Environ. Int. 8:111–116. Berglund B, Berglund U, Lindvall T (1986) Assessment of discomfort and irritation from the indoor air. In:Proceedings of IAQ-86. Managing Indoor Air for Health and Energy Conservation, American Society of Heating Refrigerating and Air Conditioning Engineers, Atlanta, USA, pp.138–149. Berglund B, Brunekreef B, Kn€oppel H, Lindvall T, Maroni M, Mølhave L (1992) Effects of indoor air pollution on human health. Indoor Air 22–25. De Bortoli A, Kn€oppel H, Pecchio E, Peil A (1986) Concentrations of selected organic pollutants in the indoor and outdoor air in northern Italy. Environ. Int. 12:343–356. (EU-ECA)Berglund B, Clausen G, De Ceaurriz J, Kettrup A, Lindvall T, Maroni M, Mølhave L, Pickering AC, Risse U, Rothweiler H, Seifert B, Younes M (1997) Total organic compounds (TVOC) in indoor air quality investigations. Report 19 of the European Collaborative Action “Indoor Air Quality & its Impact on Man”. EU-report EUR 17675 EN, Joint Research Center, Ispra, Italy.

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Frederiksson K (1979) Gifter i Bostadsluft (Indoor air pollution, in Swedish). H€alova˚rdskontakt 3:14–19. Gammage RB (1986) Volatile organic compounds. AIHA indoor environmental quality. reference manual (DE-AC05-840R-21400). Health and Safety Research Division, Oak Ridge National Laboratory, Oak ridge, TN, USA. Johannsson I (1978) Determination of organic compounds in indoor air with potential reference to air quality. Atmos. Environ. 12:1371–1377. Johansson I (1982) Kemiska Luftf€ororeningar Inomhus. En litteraturs Sammenst€allning (Chemical Indoor Air Pollution, a Swedish Review). Stockholm. Sweden: The National Institute of Environmental Medicine. Johansson I, Petterson S, Rehn T (1978) Gaschromatographic analysis of room air in recently built preschools (in Swedish). VVS J. 49:51–55. Johansson I, Petterson S, Rehn T (1979) Indoor air pollutants (in Swedish). VVS J. 50:6–7. Kjœrgaard SK, Mølhave L, Pedersen OF (1989) Human reactions to indoor air pollutants: n-decane. Environ. Int. 15:473–482. Kjœrgaard SK, Mølhave L, Pedersen OF (1991) Human reactions to a mixture of indoor air volatile organic compounds. Atmos. Environ. 24A (8):417–1426. Krause C, Mailahn W, Nagel R, Schulz C, Seifert B, Ulrich D (1987) Occurrence of volatile organic compounds in the air of 500 homes in the Federal Republic of Germany. In: Seifert B, Esdron H, Fischer M, F€uden H, Wegner J, editors. Indoor Air 87’. Berlin (FRG), Germany: Institute of Water, Soil and Air Hygiene. Lebret E, Van de Wiel HJ, Bos H, Noij D, Boleij JSM (1986) Volatile organic compounds in Dutch homes. Environ. Int. 12:323–332. Levy F (1997) Clinical features of multiple chemical sensitivity. Scand. J. Work. Environ. Health 23 (Suppl. 3):69–73. Miksch RR, Hollowell CD, Schmidt HE (1982) Trace Organic Chemical Contaminants in Office Spaces. Environ. Int. 8:129–137. Mølhave L (1986) Indoor air quality in relation to sensory irritation due to volatile organic compounds. ASHRAE Trans. 92 (1):306–316, Publication #2954. Mølhave L (1990) Volatile organic compounds, indoor air quality and health, Vol. 5. In: Walkinshaw D, editor. Indoor Air ’90: Proceedings of the 5th International Conference on Indoor Air Quality and Climate, Toronto. Ottawa, Ontario, Canada: Canada Mortgage and Housing Corporation. pp.15–33. Mølhave L (1991) Human response to volatile organic compounds as air pollutants in normal buildings. J. Exposure Anal. Environ. Epidemiol. 1(1):63–81. Mølhave L (1998) Principles for evaluation of health and comfort hazards caused by indoor air pollution. Indoor Air 8(Suppl. 4):17–25. Mølhave L, Andersen I (1980) Forureningskomponenter i indeluften i “Nulenergihuset” ved DTH (Air pollution in an experimental house, in Danish). Varme 45:121–125. Mølhave L, Møller J (1979) The atmospheric environment in modern Danish dwellings–measurements in 39 flats. In: Fanger PO, Valbjørn O, editors. Indoor Climate. Copenhagen, Denmark: Danish Building Research Institute. Mølhave L, Nielsen GD (1992) The TVOC indicator of human response to low level exposures to volatile organic compounds (VOC). Indoor Air 2:5–77. Mølhave L, Thorsen T (1991) A model for investigations of ventilation systems as sources for volatile organic compounds in indoor climate. Atmos. Environ. 25A(2):241–249. Mølhave L, Møller J, Andersen I. (1979) Luftens indhold af gasser, dampe og støv i nyere boliger (Indoor air pollution in home, in Danish). Ugesk. Lœger 141:956–961.

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Mølhave L, Andersen I, Lundqvist GR, Nielsen PA, Nielsen O (1982) Afgasning fra Byggematerialer—forekomst og Hygiejnisk Vurdering (Emission of Air Pollutants from Building Materials. In:Danish with English summary. SBI-report no. 137. Copenhagen, Denmark: The Danish Institute of Building Research. Mølhave L, Bach B, Pedersen OF (1986) Human reactions to low concentrations of volatile organic compounds. Environ. Int. 12:167–175. Mølhave L, Jensen JG, Larsen S (1991) Subjective reactions to volatile organic compounds as air pollutants. Atmos. Environ. 25A(7):1283–1293. (NAS) National Research Council (1992) Multiple Chemical Sensitivities. Addendum to Biological Markers in Immunotoxicology. Washington, DC, USA:National Research Council, National Academy Press. Otto DA, Mølhave L, Rose G, Hudnell HK, House D (1991) Neurobehavioral and sensory irritant effects of controlled exposure to a complex mixture of volatile organic compounds. Neurotoxicol. Teratol. 12:649–652. Shorter E (1997) Multiple chemical sensitivity: pseudo disease in historical perspective. Scand. J. Work Environ. Health 23(3):35–42. Wallace L (1987) The Total Assessment Methodology (TEAM) Study. Summary and Analyses, Vol. 1. Washington, DC, USA: U.S. Environmental Protection Agency. Wang TC (1975) A study of bioeffluents in a college classroom. ASHRAE Trans. 81:32–44. World Health Organization (WHO) (1961) Constitution of the World Health Organization: Basic Documents, 15th edn. Geneva, Switzerland: WHO. World Health Organization (WHO) (1982) Indoor air pollutants, exposure and health effects assessment. Euro Reports and Studies No. 78: Working Group Report. WHO Regional Office for Europe. Copenhagen, Denmark. World Health Organization (WHO) (1984) Indoor air quality research. Euro Reports and Studies No. 103. WHO Regional Office for Europe. Copenhagen, Denmark. World Health Organization (WHO) (1989) Indoor air quality: organic pollutants. Report on a WHO Meeting, Euro Reports and Studies No. 111. WHO Regional Office for Europe. Copenhagen, Denmark. Zweers T, Skov P, Valbjørn O, Mølhave L (1990) The effect of ventilation and air pollution on perceived indoor air quality in five town halls Energy Bldgs. 14:175–181.

9 FORMALDEHYDE AND OTHER ALDEHYDES George D. Leikauf

9.1 BACKGROUND Defined by a reactive, polarized carbonyl group, low-molecular-weight aldehydes are a family of organic compounds useful in a large number of industrial processes. The simplest aldehyde, formaldehyde (HCHO), is one of the top10 organic chemical feedstocks, and one of the top 20 five chemicals produced in the United States. Other widely used aldehydes include acetaldehyde (CH3CHO) and acrolein (CH2¼CHCHO), which differ from formaldehyde in carbon chain length and whether the chain is saturated or unsaturated. 9.1.1

Human Environmental Exposure

Human aldehyde exposures result from exogenous sources and endogenous formation (i.e., biogenesis through metabolism or oxidative stress) (Benedetti et al., 1980, 1984; Nilsson and Tottmar, 1987; Marnett, 1988; Anderson et al., 1997; Uchida et al., 1998; Lovell et al., 2001; Noiri et al., 2002; Shao et al., 2005). Exogenously formed or environmental aldehydes can be generated naturally through tropospheric reactions of terpenes and isoprene released from foliage with hydroxyl radicals. In addition, the major sources of aldehydes in ambient air are generated during incomplete combustion of alcohols, or are released from polymeric substances and solutions. Concern over the health effects of environmental aldehydes continues because of increasing usage of automotive fuels containing alcohols (ethanol and methanol) (Othmer, 1987). Sources other than motor vehicle exhaust include power plants, incinerators, paper mills, and refineries. This chapter reviews both the environmental sources and potential health effects, and is an update of the chapter that appeared in the second edition of this book. Other valuable reviews on aldehyde toxicity include reports by the National Research Council (NRC, 1981), Beauchamp et al. (1985), Feinman (1988), Marnett (1988),


257

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FORMALDEHYDE AND OTHER ALDEHYDES

Council on Scientific Affairs (1989), World Health Organization (WHO, 1989, 1992, 1995), Heck et al. (1990), McLaughlin (1994), International Agency for Research on Cancer (IARC, 1982, 1985, 1995a, 1995b, 2004). 9.1.1.1 Indoor Air Estimated contributions from three indoor sources of formaldehyde are listed in Table 9.1. Considering the average time–activity pattern, poor indoor air quality in the home is the most common source of aldehyde exposure. Aldehyde-generating activities in the home include tobacco smoking, wood burning, and cooking. Other common sources include release from paint, structural materials, furnishings, clothing, cosmetics, and insulation (Pickrell et al., 1983; Feinman, 1988). We typically spend less than 1 h per day (18–42 min) outside (Fig. 9.1), so that outdoor exposures constitute only 3% of our average daily exposure (Chapin, 1974; Samet et al., 1987; Samet and Spengler, 1991). Excluding the other activities, such as time in transit (1.0–1.6 h per day) and occupational exposure for those working outside the home 5.2–6.7 h per day, the remaining and primary exposure for most individuals occurs at home. Indoor exposure therefore equals 55–65% for those working outside the home and 85% for those working inside the home (Songco and Fahey, 1987). In the past, urea formaldehyde foam insulation was a source leading to the highest home exposures, including exposures averaging 120 ppb (Gupta et al., 1982). Concentrations in older homes without foam insulation typically range from 30 to 90 ppb, whereas high-level exposures of up to 4200 ppb have been recorded in mobile homes (Georghiou et al., 1983; IARC, 1985, 1995a, 1995b, 2004). A recent study of home levels determined concentrations averaged about 30 ppb (Clarisse et al., 2003; Casset et al., 2006b). Formaldehyde is also released from polymeric resins such as urea formaldehyde (used in particleboard, plywood, paper and textile treatments, and surface coatings), melamine formaldehyde (used in laminates, surface coatings, wood adhesives, and molding compounds), or phenolic resins (used in plywood adhesives and insulation) (WHO, 1989; Gerberich and Seaman, 1994). Elevation of temperature or humidity can increase release rates of formaldehyde, and thereby lead to higher exposure concentrations. More recently, penetration of outdoor ozone

TABLE 9.1

Indoor Sources of Formaldehyde Exposure

Sources

Concentration

Cigarette smoke (40 ppm in 40-mL puff) Dose per pack for smoke Environmental tobacco smoke

0.38 mg per pack 0.25 ppm

Clothing made with synthetic fibers Men’s polyester cotton blend Women’s dress

2.7 mg/g per day 3.7 mg/g per day

Furnishings Particle boarda Plywood Paneling Draperies Carpet/Upholstery Fabric

0.4–8.1 mg/g per day 1.5–5.3 mg/g per day 0.9–21.0 mg/g per day 0.8–3.0 mg/g per day 0.1 ppm

a

Made with urea–formaldehyde resin.

Source: Pickrell et al. (1983); Feinman (1988).

BACKGROUND

259

FIGURE 9.1 Time spent per day at various locations for adults. Values for working outside home are for men and women and for working in the home are for women. Adapted from Chapin (1974); and Samet et al. (1987).

and reaction with indoor materials has been found to be an additional source of aldehydes (Morrison and Nazaroff, 2002; Destaillats et al., 2002, 2006). By far, the leading indoor source of formaldehyde and several other aldehydes today is mainstream cigarette smoke (the portion inhaled by the smoker), sidestream (the portion emitted from a burning cigarette), and environmental tobacco smoke (the aged combination of sidestream and exhaled mainstream smoke) (Table 9.2). Aldehydes are principally associated with the vapor phase of mainstream smoke, but have also been measured in the particulate phase (Ayer and Yeager, 1976; Godish, 1989; Nazaroff and Singer, 2004).

TABLE 9.2

Aldehydes in Cigarette Smoke Amount Released (mg/pack) Environmental

Aldehydea Formaldehyde Acetaldehyde Acrolein Propionaldehyde a

Mainstream

Sidestream

Tobacco Smokeb

3.4 12.5 1.5 1.3

14.5 84.7 25.2 18.8

1.3 3.2 0.6 0.9

Other aldehydes in cigarette smoke include isobutylaldehyde, methacrolein, butylaldehyde, isovaleraldehyde, crotonaldehyde, and 2-methylvaleraldehyde. b Environmental tobacco smoke ¼ 2-h integrated amount per pack. Source: R. J. Reynolds (1988).

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Of the various aldehydes present in mainstream smoke, acetaldehyde is the most abundant compound by weight, followed by formaldehyde, acrolein, and propionaldehyde (Fujioka and Shibamoto, 2006). Importantly, the aldehyde concentrations released in smoke from cigarettes between puffs at a lower temperature (600 C) are greater than those produced during smoking (900 C). Thus, sidestream aldehyde emissions are 5–15 times greater than mainstream levels. Because smokers inhale only about 45% of the smoke from each cigarette, sidestream smoke can add significantly to the indoor aldehyde burden. For example, the contribution of cigarette smoke alone to indoor aldehyde concentrations can be 100 ppb formaldehyde (Gammage and Gupta, 1984; Schaller et al., 1989) or acrolein (Badre et al., 1978; Jermini et al., 1976) in rooms where several individuals are smoking. 9.1.1.2 Occupational Exposures Occupational exposure to low-molecular-weight aldehydes is extensive, primarily because of the usefulness of their reactive carbonyl in chemical synthesis. Of the several aliphatic aldehydes available, formaldehyde has the greatest production and usage. Over 10 billion pounds in the United States, and over 45 billion pounds worldwide, of formaldehyde is produced annually, with about 90% being used as a chemical feedstock or an intermediate in the synthesis of a wide number of chemicals including urea-formaldehyde and phenol-formaldehyde resins, ethylene glycol, fertili-zers, dyes, disinfectants, germicides, hardening agents, and as a preservative in water-based paints, cosmetics, and hair shampoos. Urea- and phenolformaldehyde resins are used as adhesives in the manufacture of particleboard, fiberboard, and plywood, or are used in molding, paper treating and coating, surface coating, textile treating, and insulation foam. Over one million people in the United States are estimated to be occupationally exposed to formaldehyde alone. This includes persons working in medical and health services (approximately one third), funeral homes, textiles, furniture, paper, and agriculture industries (Consensus Workshop on Formaldehyde, 1984; IARC, 1995a). Of these, over 20,000 individuals have routine exposure to concentrations greater than 1000 ppb, with over 500,000 exposed to concentrations of 500–1000 ppb (Occupational Safety and Health Administration, 1985; Noisel et al., 2007). Along with inhalation, occupational exposures to low-molecular-weight-aldehydes can also be topical, as these compounds are used as aqueous solutions (e.g., formalin, which is typically 37% formaldehyde in water and methanol) or as a polymerized solid (e.g., paraformaldehyde). Absorption through the skin is limited (typically 25% in FEV1.0

Burge et al. (1985)

Burge et al. (1985) Nordman (1985)

Frigas et al. (1984)

Alexandersson and Hedenstierna (1988) Schachter et al. (1987)

a Abbreviations. FVC: forced vital capacity; FEV1.0: forced expiratory flow at 1.0 s; FEF25–75%: mean of forced expiratory flow at 25% and 75% vital capacity; (s)Raw: (specific) airway resistance; MMEF: midmaximal expiratory flow; (s)Gaw: (specific) airway conductance.

40

2.0

Individual with Occupational Exposure 0.3–0.6 8 h work Oronasal (at work)

274


bronchial provocation. Such investigations have documented immediate and delayed responses, much like those found in antigen-induced immediate hypersensitivity reactions, but these responses appear to be rare (occurring in 12 of 230 persons tested, or 5%, in one study). Given the widespread exposure to formaldehyde and documented formaldehyde skin sensitization (see below), these results suggest that there are remarkably few cases of pulmonary hypersensitivity. This implies that sensitization by aldehyde inhalation is not a primary mechanism for a respiratory effect, but rather, observed effects are associated with a direct (nonimmuno-specific) irritant effect. This limited role for immunologic sensitization is particularly evident if hypersensitivity is narrowly defined as a response to extremely low doses of a compound, well below the dose necessary to induce irritation. Few data exist comparing possible formaldehyde effects on minute volumes in humans with this well-documented effect noted in laboratory animals. This is unfortunate, because this information would be useful if directly compared to the symptomatic responses noted within the same species. 9.2.1.4 Airway Reactivity Aldehydes can influence the underlying bronchial reactivity of the airways following an initial, transient bronchoconstriction (typically produced only at high (10,000 ppb) concentrations). Several other irritants (sulfur dioxide, ozone, and toluene diisocyanate) that induce an immediate bronchoconstriction also can induce bronchial hyperreactivity. Hyperreactivity is experimentally defined as heightened responsiveness to inhaled methacholine (a stable form of acetylcholine) or histamine, and is a diagnostic feature of asthma (Boushey et al., 1980; Barnes et al., 1989). Following a single initial exposure of healthy individuals, this condition is typically reversible and lasts for 12–48 h. In persons with asthma, however, this condition can persist for several years. Hyperreactivity frequently lacks an immunologic component, and is thus termed nonspecific. In these cases no identifiable causative antigen can be found, and a general heightened response is noted after a wide range of irritant stimuli. The relationship between each phase of the dual responses (e.g., an immediate and/or a delayed decrease in FEV1.0) after antigen presentation/challenge and hyperreactivity is currently unclear. Animal models have been informative in the past. Formaldehyde exposure of guinea pigs for 2 h produced a small change in pulmonary resistance, with an estimated half-maximal change in bronchial reactivity at 8000 ppb (Swiecichowski et al., 1993). When the duration of exposure was extended to 8.0 h, 1100 ppb formaldehyde produced a significant increase in reactivity, greater in magnitude as that produce when exposure was short but to a eightfold higher concentration. This study suggests that low-level exposures of several hours may have effects not detectable with shorter (2 h) exposures. In another animal model, ICR strain mice were sensitized intraperitoneally with dust mite allergen (Der f) prior to exposure to 0.5% formaldehyde mist once a week for 4 weeks (Sadakane et al., 2002). Mice were then challenged by intratracheal instillation with Der f and airway inflammation examined. When combined with following Der f challenge, formaldehyde exposure enhanced the histopathology (including eosinophil infiltration and goblet cell formation) and lung levels of two cytokines associated with asthma [interleukin-5 (IL-5) and chemokine (C-C motif) ligand 5 also known as Regulated on activation, normal T cell expressed, and presumably secreted (RANTES)]. Bronchial reactivity has been examined in humans and no changes have been recorded in healthy persons after exposures up to 3000 ppb formaldehyde for 3 h (Sauder et al., 1986; Kulle et al., 1987; Green et al., 1987). In studies with persons with asthma, however, Witek et al. (1987) reported a decrease in the dose of methacholine necessary to produce a 20%

SINGLE-EXPOSURE HEALTH EFFECTS

275

decrease in forced expiratory volume in 1.0 s (FEV1) in 8 of 12 individuals following a 40-min exposure to 2000 ppb formaldehyde. The mean response for all 12 subjects was not statistically different from control; thus, further study is required before any conclusions on the effect of formaldehyde on airway reactivity can be drawn. Burge et al. (1985) also has suggested that a relationship exists between bronchial responses of subjects with previous occupational exposure to formaldehyde and their underlying bronchial reactivity. Lemiere et al. (1995) reported that three individuals developed decreased FEV1 after being exposed to formaldehyde resin dust, with one of these subjects also being responsive to formaldehyde gas. Kim et al. (2001) presented a case report of an individual with occupational asthma. Working in an environment with a mean formaldehyde level of 60 ppb (with occasional peaks of 120–130 ppb), this worker had across-shift decreases in FEV1, which was reversed by inhaltion of a bronchodilatory (beta-2 adrenergic agent). In addition, shortterm exposure (20 min) at concentrations of 500 ppb in the laboratory lead to an immediate and delay (lasting over 20 h) decreased. No circulating antibody or cutaneous reactivity specific to formaldehyde–human serum albumin conjugate was detected. In another case report, by Vandenplas et al. (2004), a single accidental high level formaldehyde exposure led to changes in FEV1 and a persistent (lasting over 1 year) increase in bronchial hyperactivity to histamine, and again this seemed not to involve a specific immunological mechanism. To begin to understand the possible mechanism, Kim et al. (2002) examined human mucosal microvascular endothelial cells following in vitro exposure to formaldehyde, and found an increase in the surface expression of intercellular adehesion molecule 1 (ICAM1) and vascular cellular adehesion molecule 1 (VCAM-1). In addition, the adhesiveness between endothelial cells and eosinophils was also increased by formaldehyde exposure. This implies that formaldehyde is acting as an irritant of the nasal mucosa that may lead to an increasing the expressions of adhesion molecules and interaction that increase eosinophil trafficking (an even critical to allergic rhinitis and possible asthma). In sensitized persons with asthma, Casset et al. (2006b) found that a short, low dose preexposure (30 min 30 ppb formaldehyde) lowered the threshold dose of dust mite need to induce allergen-mediated bronchoconstriction. In addition, the late-phase reaction (defined as a 15% decrease in FEV1) was also more common in persons exposed to both allergen and formaldehyde as compared to persons exposed to allergen and filtered air. Consistent with an increased asthmatic response, the level of eosinophil cationic protein in serum or induced sputum were greater following exposure to formaldehyde with antigen. In a recent review of the literature, these authors also noted that the risk for development of asthma is increased (approximately 1.4-fold) in homes that exceed levels of 50 ppb formaldehyde (Casset et al., 2006a). Similarily, in several other studies, children exposed to formaldehyde levels of >50 ppb were at increased risk of having asthma (see Other Responses to Multiple Exposure section). 9.2.1.5 Mucociliary Clearance Formaldehyde markedly inhibits mucociliary clearance in animals and humans (Table 9.8). This effect was noted as early as 1942 by Cralley (1942), who observed an inhibition and complete stasis of ciliary beating after formaldehyde exposure. As concerns arose about the health effects of cigarette smoke, several investigators (Guillerm et al., 1961; Wynder and Hoffman, 1963; Falk, 1963; Kensler and Battista, 1963; Carson et al., 1966; Dalhamn and Rosengren, 1971; Sisson and Tuma, 1994) confirmed that aldehydes (including formaldehyde and acetaldehyde) in smoke can alter ciliary function and extended this observation, finding that formaldehyde exposure as short as 12 s could reversibly inhibit ciliary activity.

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TABLE 9.8 Concentration

Formaldehyde Inhibition of Mucociliary Clearance in Laboratory Animals Duration of Exposure

Response

Species

20 ppm

10 min

10 ppm

–

Depression of ciliary Rabbit activity without recovery Cilia stasis in 20% Rabbit

52 ppm

–

Cilia stasis in 90%

Rabbit

32 ppm

11.5 min

Cilia stasis

Guinea pig

20 ppm

4h

Rat (Sprague– Dawley)

9.6 ppm

30 min

4.4 ppm 1.4 ppm 0.2 ppm 66 mg/cm2

30 min 30 min 30 min 60 min

33 mg/cm2

60 min

16 mg/cm2

60 min

Decrease early clearance of particles; no effect on delayed clearance (alveolar) Increase mucus flow rate at 2 min; muco- and cilia stasis at 2 and 3 min Mucostasis in 8 min Increase in clearance No effect Reduced ciliary activity within 10 min Reduced ciliary activity within 10–20 min Reduced ciliary activity within 30–40 min

References Cralley (1942) Dalhamn and Rosengren (1971) Dalhamn and Rosengren (1971) Oomichi and Kita (1974) Mannix et al. (1983)

Frog palate

Morgan et al. (1984)

Frog palate Frog palate Frog palate Rabbit and pig

Morgan et al. (1984) Morgan et al. (1984) Morgan et al. (1984) Hastie et al. (1990)

Rabbit and pig

Hastie et al. (1990)

Rabbit and pig

Hastie et al. (1990)

The onset and duration of the effects of formaldehyde on mucociliary clearances are dose dependent. Formaldehyde exposures initially diminish the movement of the surface mucus layer before decreasing ciliary beat frequency (Morgan et al., 1984), a result could be due to covalent reaction of formaldehyde with mucus macromolecules that alters the physical (rheologic) properties essential for effective energy coupling with the underlying cilia. Doses of formaldehyde sufficient to inhibit ciliary activity can also reduce the number of cilia extractable from exposed tissue preparations (Hastie et al., 1990). In cilia recovered from exposed epithelia, the specific activity of ATPase and dyneins and tubulin protein content are decreased. These responses were reversible in less than 2 h after exposure, suggesting that recovery is not dependent on de novo protein synthesis, and that ciliary loss and recovery are a dynamic process. In studies with a related aldehyde of concern due to it formation from alcohol and possible interaction with cigarette smoke, Wyatt et al. (1999) found that acetaldehyde diminished ciliary beat frequency through a protein kinase C dependent pathway. In isolated tissue preparations, responses to formaldehyde are dose dependent. In the past, a single exposure to a concentration of 1000–5000 ppb was thought to have no effect, or increases mucus transport rates, whereas concentrations between 5000 and 10000 ppb depress transport (Table 9.8). However, Schafer et al. (1999) reported a decrease in ciliary beat frequency in nasal epithelial cells isolated from persons exposed to 4100 ppb formaldehyde for 2 h. Similarly, Flo´-Neyret et al. (2001) performed a dose–response analysis, and found that 1250 ppb formaldehyde inhibited mucociliary


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transport rate within 30 min. In animals, depression of overall particle clearance has been noted at a concentration of 20,000 ppb (4-h exposure) in rats (Mannix et al., 1983; Adams et al., 1987). In humans, however, nasal particle clearance was depressed at concentrations as low as 240 ppb (4–5 h), with maximal effects observed at concentrations of 400 ppb (Anderson and Molhave, 1983). These effects on clearance are most prominent in the anterior portion of the nose. Clearance is dependent on mucus composition, mucus quantity, and ciliary function. Clearance can be compromised in humans at concentrations below those necessary to produce single effects in in vitro preparations (i.e., decreased ciliary beat frequency). This suggests that this compound acts through interference with more than a single cellular or extracellular process. For example, it is probable that ciliary proteins are sensitive to formaldehyde (Hastie et al., 1990), protein–protein methylene cross-links in mucus can alter the tertiary structure of this glycoprotein (Morgan et al., 1984), or protein kinase C activation diminishes ciliary function (Salathe et al., 1993; Wong et al., 1998; Wyatt et al., 1999). Acting together, these effects could diminish particle clearance at lower doses. 9.2.1.6 Effects of Aerosol–Formaldehyde Coexposures Amdur (1960), a founder of inhalation toxicology, reported that formaldehyde-induced increases in the respiratory resistance of guinea pigs were potentiated by the simultaneous administration of submicrometer sodium chloride aerosol (sodium chloride alone produced no effect). A response, induced by a formaldehyde concentration delivered as a formaldehyde–aerosol mixture when breathed through the nose, was greater than the response to the same formaldehyde dose (gas alone) when administered directly to the lung through a tracheal cannula. This interesting observation suggests that the increment added by the aerosol is not solely a result of the transfer of more formaldehyde to the lungs but may reflect additional factors. Similarly, Kilburn and McKenzie (1978) reported that coexposure of formaldehyde with carbon particles produces greater recruitment of leukocytes into the airway epithelium and epithelial cytotoxicity (cytoplasmic vacuolization and nuclear aberrations) in trachea and bronchi of Syrian golden hamsters. This effect peaked 24–48 h after exposure. Because many environmental exposures involve coexposure to ambient and occupational aerosols, for example, resin particulate matter (Lemiere et al., 1995), future investigations clarifying these issues would add greatly to our understanding of how complex mixtures act in concert to exert aldehyde toxicity. 9.2.2

Other Aldehydes

9.2.2.1 Symptomatic Responses and Respiratory Mechanics The decrease in respiratory rate, induced by a wide array of aldehydes, has been studied in mice by Steinhagen and Barrow (1984). In this experimental system, a,b-unsaturated aliphatic aldehydes, acrolein and crotonaldehyde, decreased respiratory rate at half-maximal concentration (ED50) of 1000 and 3500 ppb as compared to 3100 ppb for formaldehyde. In contrast, the half-maximal dose for acetaldehyde was much greater (>2.85 ppm). Acrolein is consistently more potent that formaldehyde in multiple assays. Saturated aliphatic aldehydes with two or more carbons (e.g., butyraldehyde or propionaldehyde) had half-maximal concentrations of 0.75–4.2 ppm, whereas cyclic aldehydes (e.g., 3-cylcohexane-1-carboxyldehyde or benzaldehyde) exert its effects in intermediate doses ranging from 60000 to 400,000 ppb. Thus, the relative potency was acrolein > crotonaldehyde formaldehyde > benzaldehyde acetaldehyde.

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This apparent relationship holds for most other toxic endpoints including measures of pulmonary function in humans (Pattle and Cullumbine, 1956), half-maximal lethal dose in mice, guinea pigs, and rabbits (Salem and Cullumbine, 1960), and nasal pathology in rats (Lam et al., 1985; Roemer et al., 1993; Cassee et al., 1996b). Acrolein, at doses less than 1000 ppb, can produce a number of pulmonary effects. Murphy et al. (1963) reported that increases in pulmonary resistance and decreases in respiratory rate in guinea pigs exposed to 400–1000 ppb acrolein. As with exposure to formaldehyde, these changes were rapid in onset, remained somewhat constant (“response plateau”) during exposure, and reversed within 60 min after exposure. Atropine inhibited the acrolein-induced change in resistance, suggesting that involvement of a vagally mediated cholinergic pathway. That this response is so readily reversible, with cessation of exposure, also suggests that continuous occupancy of an irritant receptor by acrolein during inhalation is necessary for the initiation of this reflex. Microelectrode recordings of trigeminal nerve fibers during inhalation of aldehydes (Kulle and Cooper, 1975) are consistent with involvement of this neural pathway in the decrease in respiratory rate (Kane and Alarie, 1977). The role of specific sensory afferent pathways was further examined by Lee et al. (1992), who reported that acrolein inhalation evoked an inhibitory effect on breathing with a prolongation of expiration and bradycardia. As determined by a number of interventions, acrolein activated both vagal C-fiber active afferents and rapidly adapting irritant receptors, and suggested that the elongation of expiration was due to stimulation of the former afferent pathway. Similar to acrolein, formaldehyde stimulates C-fiber nerves and can stimulate the release of substance P, which may induce a neurogenic inflammatory response. In rat airways, this is marked by an increase in vascular permeability that is mediated predominantly by stimulation of the tachykinin NK1 receptor (Ito et al., 1996). In addition, Fujimaki et al. (2004) found that mice exposures to 2000 ppb formaldedhye for 12 weeks increased substance P levels in plasma. High acrolein concentrations (>20,000 ppb) are an irritant component of smoke, and are thought to be a causative agent in pulmonary edema, pulmonary hypertension, and acute lung injury that result from such exposures (Hales et al., 1988, 1992; Barrow et al., 1992). This process is dependent on eicosanoid formation and inhibitors of specific pathways may be benefitial therapeutic approaches (Janssens et al., 1994; Hales et al., 1995). 9.2.2.2 Airway Reactivity In addition to the transient increase in baseline pulmonary resistance, acrolein exposure of 800 ppb (2 h) can produce bronchial reactivity in guinea pigs (Leikauf et al., 1989a). Hyperreactivity occurred as early as 1 h after exposure to 1300 ppb, became maximal at 2–6 h, and lasted for longer than 24 h. This response was accompanied by an increase in three bronchoactive eicosanoids (prostaglandin F2a thromboxane B2, and leukotriene C4) in bronchoalveolar lavage fluid. Inhibition of 5-lipoxygenase diminished the response (Leikauf et al., 1989b), indicating lower respiratory tract epithelial injury. An increase in leukocyte infiltration was also noted, occurring 6–24 h after exposure. These findings suggest that acrolein-induced hyperreactivity occurred by a pathway dependent on acute injury to the airway epithelium and mediator release. In addition, migration of leukocytes into the airway does not precede hyperreactivity, suggesting that injury to cells normally present in the lung during exposure produces the mediators responsible for hyperreactivity. Ben-Jebria et al. (1994, 1995 found that lumenal exposure of isolated ferret trachea to 300 ppb acrolein for 1 h decreased the contractile dose of cholinergeric agonists (carbachol or acetylcholine), and increased the maximal contraction (indicative of increased smooth muscle reactivity). Subsequent studies with


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human and rat tracheal smooth muscle exposed to acrolein (effective dose typically 0.1–0.2 mM) demonstrated that cholinergic enhancement of contraction was accompanied by increased membrane current (Hyvelin et al., 2001), and oscillations in intracellular calcium (Roux et al., 1998; Hyvelin et al., 2000). Turner et al. (1993a) examined airway responses to intravenous substance P following acrolein exposure. Guinea pigs were exposed twice to 1600 ppb acrolein (7.5 h/day on 2 consecutive days) and followed for up to 28 days. Again, pulmonary inflammation and epithelial damagewere prominent 1 day after acrolein exposure. Neutral endopeptidase (NEP) activity was decreased in the lungs, trachea, and liver 1 and 7 days after acrolein exposure. Twenty-eight days after exposure, NEP activity in the lungs and liver was not significantly different in vehicle- and acrolein-exposed guinea pigs, but was still reduced in tracheal tissue. Acrolein increased airway reactivity to substance P that lasted for up to 28 days following exposure. Thiorphan, a NEP inhibitor, potentiated this response. To further investigate the role of neuropeptides in acrolein-induced airway responses, a subsequent study with capsaicintreated guinea pigs exposed to acrolein was performed (Turner et al., 1993b). Capsaicin depletes neuropeptides from C-sensory fibers and resulted in 100% mortality (12 of 12 guinea pigs) within 24 h of two 7.5 h 1600 ppb exposures. This compared with only 14% mortality in guinea pigs exposed to acrolein alone. Pretreatment with capsaicin also exacerbated pulmonary inflammation and epithelial necrosis and denudation. Thus, acrolein activates airway C-fibers, which release neuropeptides and alters breathing. The resulting shallow breathing patterns may be protective by reducing deposition in the distal airways. Human studies have been performed with acetaldehyde. Myou et al. (1993) found that nine subjects with asthma had a dose-dependent decrease in FEV1 following 2-min inhalations of an aerosol of saline and acetaldehyde (0, 5, 10, 20, and 40 mg/mL) solution. In control subjects without asthma, acetaldehyde had no effect on lung function. Five of the subjects with asthma also had alcohol-induced bronchoconstriction, common among the Japanese population. As stated above, alcohol sensitivity is due to a dual inheritance of a rapid isoform of alcohol dehydrogenase with a slow isoform of aldehyde dehydrogenase-2. This allelic combination increases the formation of acetaldehyde and decreases subsequent clearance, and is carried by 20% of persons of Oriental lineage. It is further enriched among Japanese asthma patients, with over 50% of these individuals having alcohol intolerance (Gong et al., 1981; Watanabe, 1991). However, this factor did not explain the effect of inhaled acetaldehyde observed by Myou et al. (1993), because four of the nine asthma patients were not alcohol intolerant, yet they responded like those patients that were. Although the sample size of these populations is small, it is possible that persons with asthma may be at greater risk to the bronchoconstrictive effects of acetaldehyde. These results are like the immediate bronchoconstriction that has been noted after sulfur dioxide exposure (Sheppard et al., 1984). 9.2.2.3 Mucociliary Clearance and Defense Mechanisms Along with studies with formaldehyde, mucociliary function has been studied with other aldehydes. Unlike reported bronchoconstriction, Dalhamn and Rosengren (1971) found that the effects and dosimetry of formaldehyde were comparable to acrolein and crotonaldehyde in potency in the inhibition of cilia beating. Acetaldehyde was much less potent. Changes in ciliary beat can also be accompanied by alteration in mucus load. Borchers et al. (1998) found that acrolein exposure increases mucin (predominately MUC5ac) transcript and protein levels in rats. Furthermore, Borchers et al. (2008) found that CD8(þ) T cells are important mediators of macrophage accumulation in the lung and the progressive airspace enlargement in response to chronic acrolein exposures. The expression of several inflammatory cytokines (IP-10, IFN-gamma,

280


IL-12, RANTES, and MCP-1), MMP2 and MMP9 gelatinase activity, and caspase3 immunoreactivity in pulmonary epithelial cells were attenuated in the Cd8-deficient mice compared to wild-type. Because acrolein has greater penetration of the upper respiratory tract than formaldehyde, effects of acrolein on other lung defense mechanisms has been examined. Exposure to 1000–2000 ppb acrolein significantly suppresses intrapulmonary killing in mice challenged with bacteria (Jakab, 1977; Astry and Jakab, 1983; Aranyi et al., 1986; Jakab, 1993), suggesting that host defense mechanisms of distal airway and alveolus are impaired. Macrophage activation and dysfunction induced by acrolein have also been observed in vitro (Leffingwell and Low, 1979; Grundfest et al., 1982; Sherwood et al., 1986; Jakab, 1993; Li et al., 1997). Thus, damage to the lower respiratory tract is much more likely after exposures to aldehydes other than formaldehyde, and some of these compounds (i.e., unsaturated aliphatic aldehydes) have equal or greater potency than formaldehyde. Even though acrolein can penetrate to the distal lung more than formaldehyde, the effects of acrolein (2500 ppb) can be enhanced by coexposure with carbon black particulate matter (10 mg/m3). Following exposure to each agent or coexposures of 4 h/day for 4 days, Jakab (1993) challenged Swiss strain mice with different infectious agents: Staphylococcus aureus to evaluate alveolar macrophage (AM) surveillance, Proteus mirabilis to evaluate AMs and polymorphonuclear leukocytes (PMNs) surveillance, Listeria monocytogenes to evaluate lymphokine-mediated cellular immunity, and influenza A virus for the cytotoxic T-cellmediated of cellular immunity. Only coexposures produced effects suppressed the intrapulmonary killing of S. aureus a day after exposure with a return to control levels by Day 7. In contrast, the coexposure enhanced the intrapulmonary killing of P. mirabilis possibly resulting from a significant increase in PMNs recovered in lung lavage fluid (also only noted following coexposure with infection). Combined exposure to carbon black and acrolein also impaired elimination of L. monocytogenes and influenza Avirus from the lungs. Exposure of alveolar macrophage to these concentrations directly suppressed Fc-receptor-mediated phagocytosis for up to 11 days (Jakab and Hemenway, 1993), which agrees with diminished S. aureus surveillance. However, the effect was short-lived in vivo. These studies suggest that the carbon black particle acts as a carrier for acrolein to enhance penetration to the distal airway and alveolar regions of the lung. The mechanisms of diminished innate immunity following acrolein exposure are not fully understood. Witz et al. (1985) noted the reactive aldehydes diminish superoxide anion production by neutrophils. In addition, a,b-unsaturated aldehydes have been found to alter NADH activity and macrophage and neutrophil membrane function, fluidity and sulfhydryl status (Witz et al., 1988). These studies were extended by Li et al. (1997) that demonstrated that acrolein increased cell death (necrosis) and programmed cell death (apoptosis) of macrophages. These effects were thought to be mediated by secondary intracellular formation of oxidants (Nardini et al., 2002; Misonou et al., 2006; Yousefipour et al., 2005) and activation of NFkB, and acrolein was also found to inhibit endotoxin-induced NFkB activation and decrease the basal level NFkB activity, which may be responsible for the inhibition of cytokine release and the induction of apoptosis in human alveolar macrophages (Li et al., 1999). The induction of apoptosis may be mediated by caspase-7 and -9 activation that induces cytochrome c release from the mitochondria (Tanel and Averill-Bates, 2005; Luo et al., 2005). In addition, acrolein and crotonaldehyde may diminish innate immunity by diminishing T-cell cytokine production, which can be reversed by the thiol compound, N-acetylcysteine (Lambert et al., 2005). Lastly, recent studies suggest that in vivo acrolein exposure of mice suppresses LPS-induced Th1 cytokine responses without affecting acute

EFFECTS OF MULTIPLE EXPOSURES

281

neutrophilia (Kasahara et al. 2008). This finding suggests that cytokine signaling can be disrupted by acrolein may represent a mechanism by which smoking contributes to chronic disease in chronic obstructive pulmonary disease and asthma. In contrast, acrolein diminishes apoptosis of neutrophils (Finkelstein et al., 2001). The inhibition of constitutive neutrophil apoptosis is mediated by common mechanisms, involving changes in cellular-reduced glutathione (GSH) status, resulting in reduced activation of initiator caspases as well as inactivation of caspase-3 by modification of its critical cysteine residue (Finkelstein et al., 2005). It is thought that diminished apoptosis in leukocytes at site of injury can lead to persistent inflammation.

9.3 EFFECTS OF MULTIPLE EXPOSURES 9.3.1

Formaldehyde

9.3.1.1 Carcinogenesis According to IARC (2004), formaldehyde is carcinogenic to humans (Group 1) on the basis of sufficient evidence in experimental animals and in human. This is a higher classification than the previous IARC evaluations of Group 2A––probably carcinogenic to humans, which were based on sufficient evidence in experimental animals, but insufficient evidence in humans (IARC, 1995a). In addition, in vitro and in vivo genotoxic studies support this classification. The major reason for the stricter classification was due to updated epidemiological studies that found that occupational formaldehyde exposure was associated with induction of nasopharyngeal cancer (Hildesheim et al., 2001; Hauptmann et al., 2004). Formaldehyde has genotoxic effects in a wide number of in vitro systems (Auerbach et al., 1977; Ma and Harris, 1988; WHO, 1989; Feron et al., 1991; IARC, 1995a). The level of activity in many experimental systems is often either weak or moderate compared to other known mutagens and can depend on what are sometimes unique experimental conditions. Regardless of these caveats, formaldehyde is clearly capable of reacting with cellular macromolecules (see above), and DNA damage induced by formaldehyde includes DNA– protein cross-links and single-strand breaks (Fornace, 1982; Fornace et al., 1982; Graftstrom et al., 1984; IARC, 1995a). DNA–protein cross-links have been found in vivo in the nasal passages of Fisher 344 rats and rhesus monkeys (Heck et al., 1989; Casanova et al., 1991). Graftstrom et al. (1984) also found that DNA single-strand breaks could accumulate in the presence of DNA excision repair inhibitors and that formaldehyde can inhibit repair. Studies examining genotoxic effects in bacteria, yeast, fungi, and Drosophila produce varied responses. For example, mutations in Drosophila are only noted in male larvae under culture conditions requiring adenosine, adenylic acid, or RNA medium supplementations. In mammalian systems the results are also subtle, complex, and varied. As noted by Boreiko and Ragan (1983), formaldehyde-induced sister-chromatid exchange in Chinese hamster ovary cells has been found by one investigator (Obe and Beck, 1979), but not by another (Brusick, 1983). In addition, formaldehyde is not mutagenic in either of these cells (Hsie et al., 1979). Another illustration of the complexity associated with the evaluation of formaldehyde’s genotoxic effects is that it can induce unscheduled DNA synthesis, which has been recorded in HeLa cells (Martin et al., 1978), but not in monkey kidney cells (Nocentini et al., 1980). In vivo dominant lethal assays in mice are negative or transiently positive (Epstein, 1972; Fontignie-Houbrechts, 1981). In any case, formaldehyde can facilitate malignant transformation of mouse embryo C3H10T1/2 C18 fibroblasts, acting

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through either an “initiating” or “promoting” mechanism (Ragan and Boreiko, 1981; Boreiko and Ragan, 1983; Frazelle et al., 1983), and produce DNA single-strand breaks and DNA–protein cross-links (Graftstrom et al., 1984). Mutational spectra induced by formaldehyde have been studied in human lymphoblasts in vitro, in Escherichia coli and in naked pSV2gpt plasmid DNA (Crosby et al., 1988). In lymphoblasts large visible deletions of some or all the X-linked hprt bands were detect by Southern blot analysis Liber et al., 1989). In E. coli, mutations in the xanthine guanine phosphoribosyl transferase (gpr) gene included large insertions (41%), large deletions (18%), and point mutations (41%). Many of the point mutations were GC transversions. Formaldehyde-induced genetic alterations in E. coli are concentration dependent, with higher doses yielding 92% point mutations, 62% of which were single AT transitions. Exposure of a SV2gpt plasmid DNA (with transformation into E. coli) resulted in frame shift mutations. Through a comparison of DNA repair proficient and deficient heterokaryons of Neurospora crassa, de Serres and Brockman (1999) examined formaldehyde induced specific-locus mutations at two closely linked loci in the adenine-3 (ad-3) region and inactivation of heterokaryotic conidia. As in many other systems, formaldehyde was a weak mutagen in the DNA repair proficient strain. However, in the DNA repair-deficient strain, formaldehyde caused about a 35-fold higher frequency of ad-3 mutations and pronounced inactivation of heterokaryotic conidia. In addition, formaldehyde induced a 5.4-fold higher frequency of ad3 mutations resulting from multilocus deletion mutation. As with X-ray-induced multilocus deletion mutations, the formaldehyde-induced ad-3 mutations could lead to deleterious heterozygous effects. Overall, the in vitro assays demonstrate the feasibility that formaldehyde could be mutagenic, but the results are complex and not always consistent. In contrast, studies with laboratory animals are clearer. Chronic inhalation of formaldehyde induces squamous cell carcinoma in the nasal cavity of rats (Swenberg et al., 1980; Albert et al., 1982; Kerns et al., 1983; Sellakumar et al., 1985; Feron et al., 1988; Woutersen et al., 1989). Exposure to 14,300 ppb formaldehyde (6 h/day 5 days/week 24 weeks) produced carcinomas in 103 of 240 (43%) Fischer 344 rats and in 2 of 240 (1%) C57BL/ 6 C3H/He F1 mice (Kerns et al., 1983; Morgan et al., 1986). (Please note the C57BL/6 mouse strain is a resistant to many lung carcinogens, and the F1 offspring of this strain is likely to be resistant). Similarly in a 28 month study (15,000 ppb formaldehyde 6 h/ day 5 days/week), Kamata et al. (1997) reported that male Fisher 344 rats developed nasal tumors that were macroscopically evident within 14 months, with 8 of 32 rats developing tumors (squamous cell papillomas and carcinomas) at 28 months. No nasal tumors were observed with lower concentrations (300 and 2000 ppb groups), but nasal epithelial cell hyperplasia, hyperkeratosis, squamous metaplasia, inflammatory cell infiltration, erosion, and edema was evident in all groups. To understand how formaldehyde exposure leads to cancer, Recio et al. (1992) examined 11 rat nasal squamous cell carcinomas and found point C or G mutations in regions II–Vof the tumor suppressor gene, p53, in five tumors. Proliferating cell nuclear antigen (PCNA) staining was similar in pattern and distribution to p53 immunoreactivity (Wolf et al., 1995). These mutations [particularly a CpG dinucleotide at rat codon 271 (codon 273 in humans)] were mutational hot spots that occur in many human cancers. Studies with different strains of rats suggest that a gene–environment interaction may be controlling the cancer response. For example, Sprague–Dawley rats also respond to 14,200 ppb formaldehyde, but the time to onset was delayed, and the total carcinoma incidence was less (38 squamous cell carcinoma (38%) and 10 polyps or papillomas in 100 rats) (Albert et al., 1982). Wistar rats yielded six tumors in 132 animals exposed to 20,000 ppb formaldehyde for 4–13 weeks and followed by observation to 126 weeks (Feron


283

et al., 1988). Although the exposure period and concentration varied, these results suggest that strain differences exist in rats (Fisher 344 Sprague–Dawley > Wistar strain) and that rats are more sensitive than one mouse strain derived from a resistant cross. Future studies to identify the genetic loci controlling these responses would be very informative. The formaldehyde concentration–carcinoma response relationship is nonlinear in rats, and also varies among strains. In Fisher 344 rats, exposure to 5600 ppb formaldehyde resulted in two rats (of 240 exposed) with squamous carcinomas, and 2000 ppb produced no carcinomas (Kerns et al., 1983). Similarly, Woutersen et al. (1989) found squamous-cell carcinoma in 1 of 26 Wistar rats exposed to 10,000 ppb for 28 months. (However, 15 of 58 rats develop carcinomas when the nasal epithelium was injured before exposure.) Swenberg et al. (1983) ascribed this nonlinearity in response to be due to nonlinearity in the formaldehyde dosimetry to its macromolecular target site. This nonlinear dosimetry results in part from variability of the breathing pattern during exposure, which influences the inhaled dose. (As noted earlier, formaldehyde deposition in mice exposed to 14,000 ppb was equivalent to that in rats exposed to 5600 ppb. Consistent with this dosimetry estimate was the incidence of carcinomas: 2/240 in mice after 14,000 ppb and in 2/240 rats after 5600 ppb). In addition to altered breathing pattern, these investigators have ascribed the nonlinearity in the dose–response relationship to the nonlinearity of cytotoxicity and overloading of protective mechanisms (including metabolic biotransformation, mucociliary clearance, and DNA repair, all of which might be altered by multiple genetic variants) induced at higher formaldehyde concentrations. In Syrian golden hamsters, 10,000 ppb formaldehyde alone (5 h/day 5 days/week for life) did not produce nasal carcinomas (Dalbey, 1982). In the exposed group, however, the mortality was marked, with only 20 of 88 animals surviving 10 weeks of treatment. In these animals, nasal epithelial cell hyperplasia or metaplasia was observed. When hamsters were exposed to formaldehyde before exposure to another carcinogen, diethylnitrosamine, the incidence of respiratory carcinomas increased. Formaldehyde (30,000 ppb) exposures were for 48 h before subcutaneous injections with 0.5 mg diethylnitrosamine (each once a week for 10 weeks). About twice as many tracheal adenomas (per tumor-bearing hamster) were observed when compared to hamsters exposed to diethylnitrosamine alone. When formaldehyde was given after diethylnitrosamine, no increase in adenomas was observed. Formaldehyde clearly produces nasal carcinomas in rats, and although mice and hamsters may be less sensitive than rats, nasal tumors have been observed in mice, and nasal hyperand metaplasia has been observed in hamsters. The nasal epithelial disruption has been investigated in several species including mice, rats, hamsters, and monkeys (Chang et al., 1983; Rusch et al., 1983; Maronpot et al., 1986; Morgan et al., 1986; Al-Abbas et al., 1986; Zwart et al., 1988; Monticello et al., 1989; St. Clair et al., 1990; Monticello et al., 1991; Bhalla et al., 1991; Cassee et al., 1996b; Cohen Hubal et al., 1997). In each species tested, the anterior nasal epithelium portion is most affected, although some species differences exist in the regional distribution of epithelial lesions. Repeated exposures to nearly comparable concentrations (about 5000–6000 ppb) produced loss of cilia and various stages of hyperplasia and squamous metaplasia of pseudostratified columnar respiratory epithelium and leukocyte infiltration. In Fischer 344 rats, these changes were most severe in the maxilloturbinate, the lateral aspect of the nasoturbinate, and the wall of the lateral meatus (Morgan et al., 1986), whereas in the rhesus monkey changes were most severe in the middle turbinate. In both species, effects noted after a single exposure become more persistent, and the percentage of nasal surface area affected progressed, when exposures are extended. The Fisher 344 appears to be more sensitve to formaldehyde-induced nasal pathology than the Brown Norway strain of rat following a single exposure (Ohtsuka et al., 1997).

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Formaldehyde also increases nasal epithelial proliferation rate. In monkeys, effects were more persistent than those observed in rats. In rats, the proliferation rate returned to control values after 3–9 days at 1000–6000 ppb (Morgan et al., 1986; Cassee et al., 1996a, 1996b). In contrast, in monkeys the proliferation rates remained elevated longer and areas of effected epitheliumweremorewidespread.InaninvitrocorrelatedstudybyTyihaketal.(2001),10 mM formaldehyde caused extensive cell damage to human colon carcinoma (HT-29) and human endothelial (HUV-EC-C) cell cultures, 1.0 mM enhanced apoptosis and reduced mitosis, and 0.1 mMenhancedcell proliferationanddecreasedapoptoticinbothcelltypesbut moresointhe tumor cells. Changes in rat epithelium were mimicked by changes in cell cycle transcript expression (Hester et al., 2003). Increases were seen in cell proliferation in the nasopharynx, trachea, and carina of the lungs of monkeys, as compared to only in the anterior nasal cavity of rats (Monticello et al., 1989). Since monkeys, like humans, breathe through both the nose and the mouth, the involvement of the lower respiratory tract also may be possible in humans. Epidemiological studies have examined the cancer mortality among persons with occupational and residential exposure to formaldehyde. This topic remains controversial, and has been reviewed extensively elsewhere (Blair et al., 1990; Feron et al., 1991; Partanen, 1993; Tarone and McLaughlin, 1995; IARC, 1995a; Collins et al., 1997; Coggon et al., 2003; Heck and Casanova, 2004; Cole and Axten, 2004; IARC, 2004; Marsh and Youk, 2004; Collins and Lineker, 2004; Binetti et al., 2006). Based on animal toxicological data, most concern has been directed at carcinomas of the nasal cavity because formaldehyde is thought to be so reactive that the target tissue must directly be exposed to inhaled formaldehyde. Evidence supporting an association between formaldehyde and nasal cancer in humans was lacking for several years. While increased relative risk for nasal cancer had been associated with formaldehyde exposure in many case-control studies (Olsen and Asneas, 1986; Hayeset al., 1986; Roush et al., 1987; Vaughan et al., 1986a, 1986b; Luce et al., 1993; West et al., 1993; Armstrong et al., 2000; Vaughan et al., 2000; Hildesheim, 2001), there were concerns about consistency in this association because it was clearly not found in two others studies (Hernberg et al., 1983; Brinton et al., 1985). In addition, this association was not always been supported by comparable increases in the standard mortality ratio in several cohort studies for nasal cancer (Marsh, 1982; Acheson et al., 1984a, 1984b; Levine et al., 1984; Stroup et al., 1986; Blair et al., 1986; Hayes et al., 1990; Gardner et al., 1993; Marsh et al., 1994). However, a re-analysis formaldehyde exposed workers by Marsh et al. (2002), when combined with the analysis by Hauptmann et al. (2004), supported associations with cancers of the upper respiratory tract and nasal cavity with formaldehyde, importantly even after adjusting for 11 potential confounding substances (Hauptmann et al., 2004). In addition, this study was supported by several case-control that were consistent with this view (Hauptmann et al., 2005). Nonetheless, the controversy has not been resolved, and reanalyses of these data suggest that uncertainities still remain (Tarone and McLaughlin, 2005; Marsh et al., 2006). In contrast to nasal cancer, positive associations of formaldehyde exposure have been noted more frequently with nasopharyngeal cancer. This association has been noted in five of seven case-control (Olsen et al., 1984; Olsen and Asneas, 1986; Vaughan et al., 1986a, 1986b; Roush et al., 1987; West et al., 1993; Armstrong et al., 2000; Vaughan et al., 2000; Hildesheim et al., 2001) and two of five cohort (Blair et al., 1986; Hayes et al., 1990; Gardner et al., 1993; Marsh et al., 1994; Andjelkovich et al., 1995) epidemiological studies. Three meta-analyses (Blair et al., 1990; Partanen, 1993; Collins et al., 1997) of the later data suggest a small to moderate increase in relative risk (1.2 to 2.1). This relationship was further re-considered by Collins et al. (1997), who proposed that underreporting (when no nasopharyngeal cancer is observed in a number of studies of small sample size) influenced


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the outcome. They found that when studies of small sample size are added to the metaanalysis, the metarelative risk for cohort studies decreases to 1.0 (with 95% confidence interval of 0.5–1.8). It is also important to consider the small number of observed cases (less than 10 nasopharyngeal cancer in total) used in these analyses. Nonetheless, the re-analysis by Hauptmann et al. (2004) also supported associations between nasopharyngeal cancer and formaldehyde exposure (including adjustment for potential confounding substances), whereas Marsh et al. (2006) challenge these conclusions. None to slight increases in risks for other cancers have been noted and include respiratory (buccal cavity, oropharynx, pharynx, and lung) (Walrath and Fraumeri, 1983; Acheson et al., 1984b; Liebling et al., 1984; Sterling and Arundel, 1985; Stayner et al., 1986; Vaughan et al., 1986; Bertazzi et al., 1986; Blair et al., 1987) and nonrespiratory sites (brain, lymphatic and hematopoietic, prostate, and skin) (Harrington and Shannon, 1975; Walrath and Fraumeri, 1984; Blair et al., 1986; Hagmar et al., 1986; Stayner et al., 1988; Hauptmann et al., 2003; Pinkertonetal.,2004;HeckandCasanova,2004).However,most oftheincreasesincancerrisk at other sites lack correlation with duration or intensity of exposure. This has been explained in part by a strong self-selection effect because of irritant responses (healthy worker effect in occupational populations), difficulties in retrospective assessments of exposure, and potential confounding factors (e.g., wood dust exposure) (Acheson et al., 1967; Olsen et al., 1984; Sterling and Weinkam, 1988; Blair and Stewart, 1990; IARC, 1995a; Collins et al., 1997; Hauptmann et al., 2003; Pinkerton et al., 2004). For example, an increased risk of nonrespiratory cancer mortality has often been associated with formaldehyde exposure in embalmers, and these individuals have exposures to complex mixtures of materials. Some argue that formaldehyde alone is unlikely to be responsible for these systemic carcinomas, givenitsrapidmetabolismatsitesofentry.Nonetheless, thepossibleassociation withleukemia (Hauptmann et al., 2003) is supported by multiple reports of increased frequency of sister chromatid exchange in peripheral lymphocytes noted of formaldehyde exposed workers (Yager et al., 1986; Suruda et al., 1993; Ying et al., 1999; Shaham et al., 2002; Ye et al., 2005). Cytology of the nasal mucosa has also been examined in formaldehyde-exposed workers (Berke, 1987; Edling et al., 1988; Holmstro¨m, 1989a, 1989b; Boysen et al., 1990). Abnormalities observed include loss of cilia, goblet cell hyperplasia, squamous metaplasia, and mild dysplasia, but, as with cancer mortality, these changes do not exhibit dose–response relationships and are confounded by coexposure to particulate matter including wood dust. Increases in squamous metaplasia in individuals living in homes with urea–formaldehyde insulation have also been inconsistent with level of exposure (Broder et al., 1988; Broder et al., 1991). In contrast, Ballarin et al. (1992) found significant levels of epithelial abnormalities in 15 subjects exposed to urea–formaldehyde glue in a plywood factory. Exposed workers had higher frequencies of micronuclei and dysplasia of nasal epithelial cells and nasal inflammation (leukocyte infiltrates) when compared to age and sex matched control subjects. Formaldehyde exposure may lead to these effects through disruption of several cellular processes. DNA–protein cross-links can arrest DNA replication and lead to the induction of other genotoxic effects such as sister chromatid exchanges in proliferating cells (Merk and Speit, 1998). Incomplete repair of DNA–protein cross-links can lead to the formation of mutations (Barker et al., 2005), which can be detected as chromosomal aberrations and micronuclei, rather than gene mutations at specific loci. These errors, in turn, can lead to larger deletions and recombinations and thereby increase micronuclei frequency (Speit and Merk, 2002). As noted above, several studies suggest increased micronuclei frequency occur in the nasal or buccal mucosa cells of formaldehyde exposed workers (Ballarin et al., 1992; Burgaz et al., 2001; Ye et al., 2005). However, the effects are not consistent across studies

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(e.g., the threshold dose may vary greatly). Speit and Schmid (2006) have suggested that this may be due to the lack of standardization of micronucleus tests, the high assay variability, and the fact that effects can be focal, and that sampling may be incomplete. In summary, formaldehyde produces a complex array of genotoxic effects in in vitro systems and consistently has been found to induce squamous carcinoma in the nasal cavity of laboratory animals (at concentrations exceeding 5000 ppb) with correlated histopathological changes (at concentrations exceeding 500 ppb). Available epidemiological evidence is conflicting, but has recently been deemed to be sufficient to support the hypothesis that formaldehyde leads to an increased risk of cancer in exposed humans (IARC, 2004). Controversy exists as to whether formaldehyde is carcinogenic to humans (Group 1), but there is little controversy that formaldehyde is probably carcinogenic to humans (Group 2A). This is because the evidence of an increased risk of nasopharyngeal cancer in humans from epidemiological studies is suggestive but controversial, and evidence for nasal cancer is limited. Evidence for cancer at other sites (including lung) is equivocal, and hard to reconcile with exposure, although recent studies suggest that leukemia may be associated with exposure. Together these findings indicate that formaldehyde is at least a Group 2A suspected (or probable) human carcinogen (IARC, 1995a; ACGIH, 2007) and exposures should be restricted accordingly. 9.3.1.2 Other Responses to Multiple Exposures Although much attention has been place on the carcinogenic potential of formaldehyde, the irritant capacity of this compound is irrefutable. Repeated formaldehyde exposure can result in eye and upper respiratory tract irritation, declines in pulmonary function, and can initiate skin sensitization. Except for skin sensitization, these effects are often readily reversible with cessation of exposure and depend more on the exposure concentration (threshold dose) than on the exposure duration (cumulative dose). Such effects may be viewed as repeated immediate responses to acute exposure rather than persistent, irreversible dysfunction produced by cumulative degradation of defense mechanisms and initiation of compensatory processes. Eye, nose, and throat irritation are the most common complaint of individuals with occupational or residential formaldehyde exposures. In occupational studies, Alexandersson and Hedenstierna (1988) found that exposures to 300–500 ppb during a work shift led to symptoms. Likewise, Horvath et al. (1988) found that across-shift responses of sore throat and burning nose increased at concentrations 400 ppb. Control responses were 3–4%, and after exposure 8–15% responded. These responses followed a dose–response relationship with 22–36% and 33–55%, responding at 2000 ppb formaldehyde. Acetaldehyde produces nasal lesions in the rat at concentrations of 0.40 ppm (6 h/day 5 days/week 4 week) (Appelman et al., 1982), and in the hamster at concentration of 40 ppm (6 h/day 5 days/week 13 week) (Kruysse et al., 1975). The nasal area most damaged with acetaldehyde was the olfactory epithelium. In hamsters, tracheal epithelial metaplasia was noted following the 4.0 ppm and a lower 1.34 ppm exposure, suggesting that the nasal passages of the hamster are less sensitive than those of the rat. Chronic acetaldehyde exposure may produce persistent changes in lung function in that in Wistar rats exposed to 243,000 ppb acetaldehyde (8 h/day 5 days/ week 5 week) had altered functional residual capacity, residual volume, total lung capacity, and respiratory frequency (Saldiva et al., 1985). Less is currently known about the potential pulmonary effects of repeated exposures to other aldehydes. Longer-chain saturated aliphatic aldehydes (e.g., propionaldehyde) are less toxic, but cyclic aldehydes (e.g., benzaldehyde) have intermediate toxic potency when compared to formaldehyde. Inasmuch as a number of environmental exposures involve the cogeneration of these aldehydes with formaldehyde, need exists for more details on the extent of environmental exposure (through routine environmental sampling) and toxicological structure–activity relationships of these compounds. One aldehyde that may need further study is glutaraldehyde becaue occupational asthma has been reported following exposure (Chan-Yeung et al., 1993; Gannon et al., 1995; Ong et al., 2004). Lastly, a number of aldehydes, including acrolein, malondialdehyde, or hexanal, are naturally occurring or produced metabolically during oxidative metabolism, which is likely to be augmented by chronic inflammation when combined with exogenous exposure to another aldehyde. A better understanding of the molecular and cellular toxicology of these and other aldehydes is warranted, and may be useful to establish a relationship between human exposure and spontaneous carcinogenesis and perhaps even chronic obstructive pulmonary disease.

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Vasiliou V, Pappa A, Estay T (2004) Role of human aldehyde dehydrogenases in endobiotic and xenobiotic metabolism. Drug Metab. Rev. 36:279–299. Vaughan TL, Strader C, Davis S, Daling S (1986) Formaldehyde and cancers of the pharynx, sinus, and nasal cavity. 1. Occupational exposures. Int. J. Cancer 38:677–683. Vaughan TL, Strader C, Davis S, Daling JR (1986b) Formaldehyde and cancers of the pharynx, sinus and nasal cavity. II. Residential exposures. Int. J. Cancer 38:685–688. Vaughan TL, Stewart PA, Teschke K, Lynch CF, Swanson GM, Lyon JL, Berwick M (2000) Occupational exposure to formaldehyde and wood dust and nasopharyngeal carcinoma. Occup. Environ. Med. 57(6):376–384. Vegheli PV, Osztovics M (1978) The alcohol syndromes: The intrarecombigenic effects of acetaldehyde. Experientia 34:195–201. Walrath J, Fraumeri Jr, JF (1983) Mortality patterns among embalmers. Int. J. Cancer 31:407–411. Walrath J, Fraumeri Jr, JF (1984) Cancer and other causes of death among embalmers. Cancer Res. 44:4638–4641. Watanabe T (1991) Mechanism of ethanol-induced bronchoconstriction in Japanese asthmatic patients. Japan J. Allergy 40:1210–1217. Watson MA, Stewart RK, Smith GBJ, Massey TE, Bell DA (1998) Human glutathione S-transferase P1 polymorphisms: relationship to lung tissue enzyme activity and population frequency. Carcinogenesis 19:275–280. Wayne LG, Bryan RJ, Ziedman KJ (1977) Irritant Effects of Industrial Chemicals: Formaldehyde.US DHEW, PHS:NIOSH Publication. 77:117. Weber-Tschopp A, Fisher T, Grahdjean E (1977) Reizwirkugen tea Formaldehyds (HCHO) auf den Menschen. Int. Arch. Occup. Environ. Health 39:207–218. Wenzlaff AS, Cote ML, Bock CH, Land SJ, Schwartz AG (2005) GSTM1, GSTT1 and GSTP1 polymorphisms, environmental tobacco smoke exposure and risk of lung cancer among never smokers: a population-based study. Carcinogenesis 26(2):395–401. West S, Hildesheim A, Dosemeci M (1993) Non-viral risk factors for nasopharyngeal carcinoma in the Philippines: results from a case-control study. Int. J. Cancer 55(5):722–727. WHO. World Health Organization. International Programme on Chemical Safety. (1989) Environmental Health Criteria 89. Formaldehyde. Geneva:World Health Organization. WHO. World Health Organization. International Programme on Chemical Safety. (1992) Environmental Health Criteria 127. Acrolein. Geneva:World Health Organization. WHO. World Health Organization. International Programme on Chemical Safety. (1995) Environmental Health Criteria 167. Acetaldehyde.Geneva:World Health Organization. Wieslander G, Norbackvd D, Bjornsson E, Janson C, Borman G (1997) Asthma and the indoor environment: the significance of emission of formaldehyde and volatile organic compounds from newly painted indoor surfaces. Int. Arch. Occup. Environ. Health. 69:115–124. Wilhelmsson B, Holmstro¨m M (1987) Positive formaldehyde-rast after prolonged formaldehyde exposure by inhalation. Lancet 2:164. Wilkins RJ, Macleod HD (1976) Formaldehyde induced DNA-protein cross links in E. coli. Mutat. Res. 36:11–16. Windholz M (1983) Merck Index. Rahway, NJ:Merck and Co. WitekJrTJ, Schachter EN, Tosun T, Beck GJ, Leaderer BP (1987) An evaluation of respiratory effects following exposure to 2.0 ppm formaldehyde in asthmatics: Lung function, symptoms, and airway reactivity. Arch. Environ. Health 42:230–237. Witz G, Lawrie NJ, Amoruso MA, Goldstein BD (1985) Inhibition by reactive aldehydes of superoxide anion radical production in stimulated human neutrophils. Chem. Biol. Interact. 53 (1–2):13–23.

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10 AMBIENT AIR PARTICULATE MATTER Morton Lippmann

A broad variety of processes produce suspended particulate matter (PM) in the ambient air in which we live and breathe, and there is an extensive body of epidemiological literature that demonstrates that there are statistically significant associations between the concentrations of airborne PM and the rates of mortality and morbidity in human populations. The PM concentrations have almost always been expressed in terms of mass, although recent studies suggest that number concentration may correlate better with some effects than does mass (Peters et al., 1997; Stolzel et al., 2003). Also, in studies that reported on associations between health effects and more than one mass concentration, the strength of the association generally improves as one goes from total suspended particulate matter (TSP) to thoracic particulate matter, a.k.a. PM less than 10 mm in aerodynamic diameter (PM10), to fine particulate matter, a.k.a. PM less than 2.5 mm in aerodynamic diameter (PM2.5). The influence of a sampling system inlet on the sample mass collected is illustrated in Fig. 10.1. The PM2.5 distinction, while nominally based on particle size, is a means of measuring the total gravimetric concentration of several specific chemically distinctive classes of fine particles that are emitted into or formed within the ambient air as very small particles. In the former category (emitted) are carbonaceous particles in wood smoke and diesel engine exhaust. In the latter category (formed) are carbonaceous particles formed during the photochemical reaction sequence that also leads to ozone formation, as well as acidic sulfur and nitrogen oxide compounds resulting from the oxidation of sulfur dioxide and nitrogen oxide vapors released during fuel combustion, and were chemically changed by their neutralization by ammonia. The coarse particle fraction, that is, those particles with aerodynamic diameters larger than 2.5 mm, are largely composed of soil and mineral ash that are mechanically dispersed into the air. Both the fine and coarse fractions are complex mixtures in a chemical sense. To the extent that they are in equilibrium in the ambient air, it is a dynamic equilibrium in which they enter the air at about the same rate as they are removed. In dry weather, the concentrations of coarse particles are balanced between dispersion into the air, mixing


317

318

AMBIENT AIR PARTICULATE MATTER

∆MASS/∆(log Da), (µg/m-3)

70 60

Fine mode particles

Coarse mode particles TSP Hi Vol

50

WRAC

40

PM10

30 20 10 PM2.5

0 5 10 20 50 100 0.1 0.2 0.5 1.0 2 Aerodynamic particle diameter (Da), (µm) Total suspended particulate matter (TSP)

PM10 PM2.5

PM(10-2.5)

FIGURE 10.1 Representative bimodal mass distribution as a function of aerodynamic particle diameter for Phoenix, AZ, Showing effect of size-selective sampling inlet on mass collected for (a) wide-ranging aerosol classifier (WRAC), (b) standard total suspended particulate (TSP) highvolume sampler, (c) Sampler following EPA’s (PM10) criteria for thoracic dust, and (d) sampler following EPA’s criteria for fine particulate matter (PM2.5). Source: U.S. EPA (1996a, 1996b).

with air masses, and gravitational fallout, while the concentrations of fine particles are determined by rates of formation, chemical transformation, and meteorological factors. PM concentrations of both fine and coarse PM are effectively depleted by rainout and washout associated with rain. The coarse particle fraction includes those less than 10 mm, which can penetrate into the thorax and cause some of the health effects associated with PM in ambient air (PM10–2.5). Further elaboration of these distinctions is provided in Table 10.1. In the absence of any detailed understanding of the specific chemical components responsible for the health effects associated with exposures to ambient air PM10, PM2.5, and PM10–2.5, and in the presence of a large and consistent bodies of epidemiological evidence associating ambient air PM size fractions with mortality and morbidity that cannot be explained by potential confounders such as other pollutants, aeroallergens, or ambient temperature or humidity, public health authorities have established ambient air standards based on mass concentrations within the fine and coarse thoracic size fractions. This chapter summarizes the nature and extent of the health effects associated with gravimetric ambient air PM concentrations and a limited number of specific components such as sulfate and nickel. Further discussions of the health effects of some of the specific constituents of the ambient air PM are discussed in the chapters on asbestos and other mineral fibers, diesel engine exhaust, lead, nitrogen oxides, and sulfur oxides.

10.1 SOURCES AND PATHWAYS FOR HUMAN EXPOSURE As indicated in Table 10.1, fine and coarse particles generally have distinct sources and formation mechanisms, although there may be some overlap. Primary fine particles are

319

Sources:

Solubility:

Composed of:

Formed by:

Metals: compounds of Pb, Cd, V, Ni, Cu, Zn, Mn, Fe, etc.

Organic compounds with very low saturation vapor pressure at ambient temperature

Probably less soluble than accumulation mode Combustion

Largely soluble, hygroscopic, and deliquescent Combustion of coal, oil, gasoline, diesel fuel, wood

Particle-bound water

Large variety of organic compounds

Metal compounds

Sulfate

Elemental carbon

Condensation

Break-up of large solids/droplets

Accumulation

Coagulation Reactions of gases in or on particles Evaporation of fog and cloud droplets in which gases have dissolved and reacted Sulfate, nitrate, ammonium, and hydrogen ions Elemental carbon

Condensation Coagulation

Combustion, high temperature processes, and atmospheric reactions Nucleation

Ultrafine

Fine Coarse

(continued)

Resuspension of industrial dust and soil tracked onto roads and streets

CaCO3, CaSO4, NaCl, sea salt Pollen, mold, fungal spores Plant and animal fragments Tire, brake pad, and road wear debris Largely insoluble and nonhygroscopic

Fly ash from uncontrolled combustion of coal, oil, and wood Nitrates/chlorides/sulfates from HNO3/HCl/SO2 reactions with coarse particles Oxides of crustal elements (Si, Al, Ti, Fe)

Suspended soil or street dust

Mechanical disruption (crushing, grinding, abrasion of surfaces) Evaporation of sprays Suspension of dusts Reactions of gases in or on particles

Comparison of Ambient Particles, Fine Particles (Ultrafine Plus Accumulation Mode) and Coarse Particles

Formation processes:

TABLE 10.1

320

Minutes–hours Grows into accumulation mode Diffuses to raindrops 5 mm and an aspect ratio >3 (ACGIH-AIHA Aerosol Hazards Evaluation Committee, 1975). The phasecontrast optical method (PCOM) is specified in the Occupational Safety and Health Administration (OSHA) occupational health standard for asbestos. Table 12.2 summarizes

TABLE 12.2 Group

Recommended Air Concentration Limits and Standards for Asbestos Year

ACGIH ACGIH ACGIH OSHA OSHA NIOSH ACGIH

1946 1968a 1970,a 1974c 1972 1976 1976 1978,a 1980c

ACGIH OSHA OSHA OSHA

1997a 1976 1986 1992

a

Limit 6

3

5 10 particles/ft 12 fibersb/mL or 2 106 particles/ft3 5 fibers/mL 5 fibers/mL 2 fibers/mL 0.1 fiber/mL 0.2 fiber/mL for crocidolite 0.5 fiber/mL for amosite 2.0 fiber/mL for chrysotile and other forms 0.1 fiber/mL for all forms 2.0 fiber/mL 0.2 fiber/mL 0.1 fiber/mL

Notice of intent. All fiber limits based on phase-contrast optical determination at 400–450 magnification. c Adopted as threshold limit value (TLV). b

400

ASBESTOS AND OTHER MINERAL AND VITREOUS FIBERS

recommended occupational exposure limits and standards used in the United States. Detailed guidance on the use of the PCOM method has been provided by the World Health Organization (WHO, 1997). The third type of concentration index is based on the mass concentration of asbestos or on the mass concentration passing a precollector meeting the British Medical Research Council (BMRC) or American Conference of Governmental Industrial Hygienists (ACGIH) sampler acceptance criteria for “respirable” dust. Some of the recent animal inhalation studies report the chamber concentrations in terms of the “respirable” mass based on samples collected using samplers that meet the BMRC criteria. Environmental exposures have been measured either in terms of fiber count or fiber mass. Fiber counts have been made using both PCOM and electron microscopy. The reported concentrations have differed according to the size distributions of the fibers, the resolving power of the microscope, and whether there was any discrimination in the analyses according to fiber type. The fiber mass index was developed by Selikoff et al. (1972). The fibers and fiber bundles in the sample are mechanically reduced to individual fibers and fibrils, which are then identified and measured by TEM. Mass concentrations in nanograms per cubic meter are calculated from the numbers of fibers and fibrils, and their dimensions. The use of these various exposure indices has sometimes led to the development of site- or industry-specific exposure–response relationships for one or more of the asbestos-related diseases, but it has not been possible to develop any generic relationships demonstrating their more general adequacy as indices of exposure or health risk. However, as discussed later in this chapter, the landmark report of the Health Effects Institute-Asbestos Research (HEI-AR, 1991), established that: (1) short asbestos fibers (i.e., those shorter than 5 mm in length) pose little, if any, health risk; and (2) the standard analyses, that count all asbestos fibers or the mass that they represent are seldom adequate to define the concentrations of the longer fibers that do pose cancer risks. Support for the need for such a fiber-length cut-off was provided by an Expert Panel for the Agency for Toxic Substances and Disease Registry (ATSDR) (ERG, 2003a). Their report was titled: “Report on the Expert Panel on Health Effects of Asbestos and Synthetic Vitreous Fibers: The Influence of Fiber Length,” and stated: “Many of the short fibers that reach the gas-exchange region of the lung are cleared by alveolar macrophages, and the rate of clearance by phagocytosis has been found to vary with fiber length. There is a strong weight of evidence that asbestos and SVFs shorter than 5 mm are unlikely to cause cancer in humans.” Similarly, in a Peer Consultation report prepared for the US EPA, it was stated that there was agreement among the panelists convened that “the available data suggest that the risk for fibers less than 5 mm in length is very low and could be zero” (ERG, 2003b). Independent analyses of published data from chronic rat inhalation studies having fiber length, diameter, and compositional data and biological outcomes have provided key evidence that the health risks are due to long fibers, especially those longer than 10 mm (Lippmann, 1988; Berman et al., 1995; Berman and Crump, 2001). 12.2.2

Exposure Levels

Esmen and Erdal (1990) reviewed published data on human occupational and nonoccupational exposure to fibers. They concluded that for the traditionally defined asbestos fibers, that is, fibers >5 mm long, large amounts of the available data suffer from the diversity of sample collection and analysis methods. Simple generalizations suggest that occupational exposures are generally several orders of magnitude higher than environmental exposures; and currently extant data and the current routine measurement practices present significant

EXPOSURES TO FIBERS

401

difficulties in the consistent interpretation of the data with respect to health effects. The human exposure data to many nonasbestos minerals that exist in fibrous habit are very scanty, and in view of the biological activity of some of these fibers, this lack of relevant data limits interpretability. With respect to asbestos exposures in buildings, the Literature Review Committee of the Health Effects Institute-Asbestos Research (HEI-AR, 1991) grouped building occupants into three main categories with respect to ACM: C1: Bystanders or nonoccupationally exposed building occupants, for example, office workers, visitors, students, and teachers. C2: Housekeeping or custodial employees who may disturb materials in the course of routine cleaning and service functions. C3: Maintenance or skilled workers that may disturb ACM in the course of making repairs, installing new equipment, or during minor renovation activity.Two other categories not often dealt with in the context of building occupancy were identified as: C4: Abatement workers or others involved in the removal or renovation of structures with ACM. C5: Firefighters and other emergency personnel who may be present during or after the fabric of the building has been extensively damaged by fire, wind, water, or earthquake. Building employees and contractor employees may disturb ACM during the course of their normal work assignments, especially during maintenance and custodial activities. They may or may not know that they are disturbing ACM, and, if they do, they may or may not have the equipment and motivation to take the appropriate precautions to minimize their exposure to airborne fibers. Exposures of such workers can be high, and warrant special concern. By far the most numerous are the C1 building occupants. Persons in categories C2–C5 fall under OSHA regulations for personal monitoring if their exposures exceed the permissible exposure limit (PEL) of 0.1 fiber/mL (f/mL) >5 mm in length as an 8-h time-weighted average, or the OSHA excursion limit of 1 f/mL in a half-hour, respectively, both as determined by PCOM. Persons in category C1 are not covered by any federal exposure limits. There are relatively few published data on the concentrations of airborne fibers in public buildings. The HEI-AR sponsored Literature Review Panel compiled the available data, both published and such unpublished data as it could assemble (HEI-AR, 1991). In their report, the panel concluded that: A large number of buildings in the U.S. and other countries have been examined for airborne asbestos fibers within the previous 20 years, and yielded many thousands of air measurements (most unpublished). However, few building environments have been characterized in sufficient detail or sampled with sufficient analytical sensitivity to adequately describe the exposures of C1 occupants. Specific details are especially lacking for episodic and point-source releases of fibers into the air of buildings from maintenance and engineering activities, from repair and renovation operations, and from normal custodial functions. Such data as are now available on the airborne concentrations of asbestos fibers of the dimensions most relevant to human health (i.e., fibers >5 mm long) generally show average concentrations on the order of 0.00001 f/mL for outdoor rural air (except near asbestoscontaining rock outcroppings) and average concentrations up to about 10-fold higher in the outdoor air of urban environments. However, outdoor urban average concentrations above

402


FIGURE 12.1 Summary of building average airborne asbestos fiber concentrations (in fibers per milliliter, with lengths >5 mm) in public and commercial buildings. (From data as reported by HEI-AR, 1991.)

0.0001 f/mL have been reported in certain circumstances as a result of local sources; for example, downwind from, or close to, frequent vehicle braking or activities involving the demolition or spray application of asbestos products. Data on ambient indoor levels of asbestos from direct TEM measurements were averaged for each of a number of individual buildings. The following data were based on 1377 air samples obtained in 197 different buildings not involved in litigation. The overall means of the studies on these buildings ranged from 0.00004 to 0.00063 f/mL, with upper 90th percentiles ranging from 0.00002 to 0.0008 f/mL. Grouped by building category, the mean concentrations were 0.00051, 0.00019, and 0.00021 f/mL in schools, residences, and public and commercial buildings, respectively, with upper 90th percentiles of 0.0016, 0.0005, and 0.0004, respectively (see Fig. 12.1).

12.3 FIBER DEPOSITION IN THE RESPIRATORY TRACT There are five mechanisms that are important with respect to the deposition of fibers in respiratory tract airways. These are impaction, sedimentation, interception, electrostatic precipitation, and diffusion (see Fig. 1.8). Impaction and sedimentation probabilities are governed by the aerodynamic diameter of the fibers, which, for long mineral fibers, are close to three times their physical diameters (St€ ober et al., 1970; Timbrell, 1972). Most impaction occurs downstream of air jets in the larger airways, where the flow velocities are high and the momentum of a fiber propels it out of the bending flow streamlines and onto relatively small portions of the epithelial surfaces (Balashazy et al., 2005; Su and Cheng, 2006). Sedimentation, on the contrary, is favored by low flow velocity, long residence times, and small airway size. Electrostatic precipitation occurs primarily by image forces, in which charged particles induce opposite changes on airway surfaces. It is dependent on the ratio of electrical charge

FIBER DEPOSITION IN THE RESPIRATORY TRACT

403

to aerodynamic drag. Little is known about the charge levels on SVFs in the workplace. Jones et al. (1983) have shown that asbestos-fiber processing operations do generate fibrous aerosols with relatively high charge levels, and that these charge levels are sufficient to cause an enhancement of fiber deposition in the lungs. Such an enhancement of fiber deposition for chrysotile asbestos was seen in rats that were exposed by inhalation (Davis, 1976). Interception increases with fiber length. The greater the length, the more likely it is that the position of a fiber end will cause it to touch a surface that the center of mass would have missed. Diffusional displacement results from collisions between air molecules and airborne fibers. For compact particles, diffusion becomes an important deposition mechanism for diameters smaller than about 0.5 mm. Fibers of similar diameter would be more massive and therefore be displaced less by a single molecular impact. Long fibers may have nearly simultaneous impacts from several gas molecules, and their random trajectories may tend to damp the net displacement. On the contrary, a single collision near a fiber end may rotate the fiber sufficiently to alter its interception probability. The role of diffusion in fiber deposition is poorly understood. Gentry et al. (1983) measured the diffusion coefficients of chrysotile and crocidolite asbestos fibers and found good agreement with theoretical predictions for chrysotile (0.4 mm mean diameter) but poor agreement with the more rod-like crocidolite (0.3 mm mean diameter). The conductive airway region of the human lung consists of a series of bifurcating airways. The trachea is the only airway segment with a length-to-diameter ratio much greater than three. Single symmetrical fibers suspended in a laminar flow stream tend to become aligned with the flow axis as they move through a lung airway. On the contrary, fiber agglomerates or nonfibrous particles would have more random orientations that would depend on their distributions of masses and drag forces. A fiber whose flow orientation differs from axial alignment would have an enhanced probability of deposition by interception. A fiber’s alignment is radically altered as it enters a daughter airway, and this loss of alignment with the flow at the entry contributes to its deposition by interception at or near the carinal edge. To the extent that a fiber is entrained in the secondary flow streams that form at bifurcations, its deposition probability by interception is further enhanced. Sussman et al. (1991a) performed an experimental study of fiber deposition within the larger tracheobronchial airways of the human lung using replicate hollow airway casts. For crocidolite fibers with diameters primarily in the 0.5–0.8 mm range, interception increased total deposition, with the effect increasing with fiber length, especially for fibers >10 mm in length. The effect was more pronounced at 60 L/min than at 15 L/min. This is consistent with greater axial alignment of the fibers during laminar flow within the airway. Morgan and Holmes (1984) and Morgan et al. (1980) exposed rats for several hours by inhalation (nose only) to glass fibers 1.5 mm in diameter and 5, 10, 30, or 60 mm long. For fibers longer than 10 mm, essentially all were deposited, mostly in the head. These results, together with the results of their earlier studies on asbestos fibers, indicate that penetrability of airborne fibers into the rat lung drops sharply with aerodynamic diameter above 2 mm. The results reported by Morgan and Holmes provided experimental verification that increasing fiber length increases lung deposition within the tracheobronchial airways. Sussman et al. (1991b) found that the deposition patterns of fibers in the larger lung airways are similar to those for particles of more compact shapes. In other words, the added deposition due to interception increased the deposition efficiencies without changing the pattern of deposition.

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Most of the studies on particle deposition patterns and efficiencies in hollow bronchial airway casts of the larynx and the larger conductive airways of the human bronchial tree have been focused on deposition during constant flow inspirations. For studies of deposition during cyclic inspiratory flows, Gurman et al. (1984a, 1984b) used a variable-orifice mechanical larynx model (Gurman et al., 1980) at the inlet in place of the fixedorifice laryngeal models used in the prior constant flow tests. In one series of tests, two replicate casts were connected in tandem. The corresponding terminal endings were connected with rubber tubing. Deposition in the downstream cast was analyzed to determine the deposition pattern and efficiencies during expiratory flow (Schlesinger et al., 1983). Concern about sites of enhanced surface deposition density is stimulated by the observation that the larger bronchial airway bifurcations, which are favored sites for deposition, are also the sites most frequently reported as primary sites for bronchial cancer (Schlesinger and Lippmann, 1978). Deposition patterns within the nonciliated airways distal to the terminal bronchioles may also be quite nonuniform. Brody et al. (1981) studied the deposition of chrysotile asbestos in lung peripheral airways of rats exposed for 1 h to 4.3 mg/m3 of respirable chrysotile. The animals were killed in groups of 3 at 0, 5, and 24 h and at 4 and 8 days after the end of the exposure. The pattern of retention on the epithelial surfaces was examined by scanning electron microscopy of lung sections cut to reveal terminal bronchiolar surfaces and adjacent airspaces. The rat does not have recognizable respiratory bronchioles, and the airways distal to the terminal bronchioles are the alveolar ducts. In rats killed immediately after exposure, asbestos fibers were rarely seen in alveolar spaces or on alveolar duct surfaces except at alveolar duct bifurcations. There were relatively high concentrations on bifurcations nearest the terminal bronchioles and lesser concentrations on more distal duct bifurcations. In rats killed at 5 h, the patterns were similar, but the concentrations were reduced. Subsequent studies have shown that crocidolite asbestos (Roggli et al., 1987), Kevlar aramid synthetic fibers (Lee et al., 1983), and particles of more compact shape (Brody and Roe, 1983) deposit in similar patterns, and that the deposition patterns seen in the rat also occur in mice, hamsters, and guinea pigs (Warheit and Hartsky, 1990). The sudden enlargement in air path cross section at the junction of the terminal bronchiole and alveolar duct may play a role in the relatively high deposition efficiency at the first alveolar duct bifurcation. Little has previously been known about the flow profiles in this region of the lung. However, Briant (1988) has shown that a net axial core flow in a distal direction and a corresponding net annular flow in a proximal direction take place during steady-state cyclic flow in tracheobronchial airways and that this could account for such concentrated deposition on the bifurcations of distal lung airways.

12.4 FIBER RETENTION, TRANSLOCATION, DISINTEGRATION, AND DISSOLUTION The fate of fibers deposited on surfaces within the lungs depends on both the sites of deposition and the characteristics of the fibers. Within the first day, most fibers deposited on the tracheobronchial airways are carried proximally on the surface of the mucus to the larynx, to be swallowed and passed into the gastrointestinal tract. The residence time for fibers on the surface of the tracheobronchial region is too short for any significant change in the size or composition of the fibers to take place.

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Fibers deposited in the nonciliated airspaces beyond the terminal bronchioles are more slowly cleared from their deposition sites by a variety of mechanisms and pathways. These can be classified into two broad categories, that is, translocation and disintegration. 12.4.1

Translocation

Translocation refers to the movement of the intact fiber: (1) along the epithelial surface to dust foci at the respiratory bronchioles; (2) onto the ciliated epithelium at the terminal bronchioles; or (3) into and through the epithelium, with subsequent migration to interstitial storage sites within the lung, along lymphatic drainage pathways, and for very thin short fibers, access via capillary blood to distant sites, as suggested by Monchaux et al. (1982). Boutin et al. (1996) suggested that thin fibers longer than 5 or 10 mm migrate toward the parietal pleura via the lymphatic pathway, where they accumulate preferentially in anthracotic “black spots” of the parietal pleura. In a study by Dodson et al. (1990) comparing the fiber content of tissues from chronically exposed shipyard workers, they reported that while 10% of amphibole fibers in pleural plaque samples were longer than 5 mm and 8% were longer than 10 mm, the corresponding figures for chrysotile fibers were 3.1 and 0%. In lymph nodes, the corresponding figures for >10 and >5 mm lengths were 6.0 and 2.5% for amphiboles and 0 and 0% for chrysotile. In lung tissue, they were 41.0 and 20.0% for amphiboles and 14.0 and 4.0% for chrysotile. Boutin et al. (1996) noted that the black spots that concentrate longer fibers were in close contact with early pleural plaques. These studies indicate that fiber translocation is dependent on both fiber diameter and fiber length, and that length is an important determinant of biological responses. Translocation may also occur after ingestion of the fibers by alveolar macrophages if the fibers are short enough to be fully ingested by the macrophages. Holt (1982) proposed that fibers phagocytosed by alveolar macrophages are carried by them toward the lung periphery by passing through alveolar walls and that some of these cells aggregate in alveoli near larger bronchioles and then penetrate the bronchiolar wall. Once in the bronchiolar lumen, they can be cleared by mucociliary transport. Lentz et al. (2003) reviewed the literature on the dimensions of fibers that may translocate to the parietal pleura, and concluded that the critical dimensions were: diameter 25 mm), the disappearance was independent of length and lung burden, implying that the clearance of such fibers occurs by dissolution and fragmentation into shorter lengths. 12.4.4

Role of Fiber Dissolution

A differential lung clearance between fibers of chrysotile and the more rod-like amphibole asbestos fibers was shown for rats that underwent chronic inhalation exposures (Wagner et al., 1974). The lung fiber burdens of the amphiboles rose continuously throughout 2 years of exposure, and declined slowly in the rats removed from exposure after six months. By

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contrast, the lung burdens in rats exposed to both Quebec and Zimbabwe chrysotile rose much more slowly during exposure, and seemed to decline after 12 months, even with further exposure. The biopersistence of chrysotile fibers from other locations has been studied by Bernstein and colleagues following inhalation exposures to aerosols with large number concentrations of fibers >20 mm in length. For chrysotile from the Cana Brava mine in central Brazil, the clearance half-time of the fibers >20, those 5–20, and those 20, those 5–20, and those 20 mm in length was 11.4 d, which was similar to that for glass and stone wools previously studied (Bernstein et al., 2005b) Similar differential retention has been found in humans. Churg (1994) reported on analyses of lung tissue for 94 chrysotile asbestos miners and millers from the Thetford region of Quebec, Canada. The retained chrysotile and exposure atmosphere contained a very small percentage of tremolite, yet the lungs contained more tremolite than chrysotile, and the tremolite content increased rapidly with the duration of exposure. While most of the inhaled chrysotile was rapidly cleared from the lungs, a small fraction seemed to be retained indefinitely. After exposure ended, there was little or no clearance of either chrysotile or tremolite from the lungs. Albin et al. (1994) studied retention patterns in lung tissues from 69 Swedish asbestoscement workers and 96 controls. They reported that chrysotile has a relatively rapid turnover in human lungs, whereas amphiboles (tremolite and crocidolite) have a slower turnover. They also noted that: (1) chrysotile retention may be dependent on dose rate; (2) chrysotile and crocidolite retention may be increased by smoking; and (3) that chrysotile and tremolite retention may be increased by the presence of lung fibrosis. The most direct evidence for the effect of altered dust clearance rates on the retention of inhaled fibers in humans comes from studies of the fiber content of the lungs of asbestos workers in various countries. Timbrell (1982) developed a model for fiber deposition and clearance in human lungs based on his analysis of the bivariate diameter and length distributions found in air and lung samples collected at an anthophyllite mine at Paakkila in Finland. The length and diameter distributions of the airborne dust at this particular mine were exceptionally broad, and historic exposures were very high. For workers with the highest exposure and most severe lung fibrosis (Ashcroft et al., 1988), the fiber distributions in some tissue segments approached those of the airborne fibers. Adjacent tissue, analyzed for extent of fibrosis, showed severe fibrotic lesions. He concluded that long-term retention was essentially equal to deposition in such segments. Fig. 12.2 shows a series of retention curves for different degrees of lung fibrosis. These curves were determined by comparing the anthophyllite fiber size distributions in other tissue samples from the same lung with the distribution in the sample for which all fibers deposited were retained. Lung fibrosis is associated with increased fiber retention, and fiber retention is clearly associated with fiber length and diameter. The critical fiber length for mechanical clearance from the lungs is 17 mm. More precise descriptions of the effect of fiber loading in the lung on fibrosis need to be based on the use of the most appropriate index of fiber loading.

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FIGURE 12.2 Effect of lung fibrosis on fiber retention in human lungs as a function of fiber length. The scores for fibrosis are A: minimal, B: slight, C: moderate, D: marked, and E: severe. (Source: Lippmann and Timbrell, 1990.)

Morgan et al. (1982) and Morgan and Holmes (1984) studied the retention of 1.5 mm diameter glass fibers administered to rat lungs by intratracheal instillation. Retention at 1 year for 5 mm long fibers was 10%, whereas for 10 mm long fibers it was 20%. For the fibers that were 30 or 60 mm long, there was no measurable clearance during the first 9 months. Further retention measurements were not made for these long fibers because of evidence that they were disintegrating and dissolving. This macrophage-mediated mechanical clearance is less effective for 10 mm long fibers than for 5 mm fibers, and is ineffective for fibers 30 mm in length and longer. The results were based on the sizes of fibers recovered from rats’ lungs at various times following inhalation exposures. For the glass fibers, there was much less dissolution of the 5 and 10 mm fibers than of the 30 and 60 mm fibers. The dissolution of the long 1.5 mm diameter fibers was very nonuniform. Some were little changed in dimension, whereas others were reduced in diameter to 0.2 mm. On the contrary, for rockwool fibers >20 mm in length, there was no observable change in fiber dimensions after 6 months. Morgan and Holmes (1984) attributed the dependence of dissolution on fiber length to the differences and intra- and extracellular pH. The shorter fibers within macrophages are exposed to a pH of 7.2, whereas those outside were exposed to the extracellular pH of 7.4. Bernstein et al. (1984) and Hammad (1984) also found evidence of substantial in vivo dissolution of glass fibers. LeBouffant et al. (1984) used X-ray analysis on individual fibers recovered from lung tissue to show the exchange of cations between the fibers and the tissues. For example, the fibers can lose calcium and gain potassium. Insight on the solubility of fibers in vivo has also been obtained from in vitro solubility tests. Griffis et al. (1981) found that glass fibers suspended either in buffered saline or serum simulant at 37 C for 60 days exhibited some solubility and that the sodium content of the residual fiber was reduced. Forster (1984) used Gamble’s solution for tests on samples of

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18 different SVFs at temperatures of 20 and 37 C and for exposure times ranging from 1 h to 180 days using static tests, tests with once-daily shaking, tests with continuous shaking, and tests with single fibers in an open bath. There was some solubility for all fibers. Klingholz and Steinkopf (1982, 1984) studied dissolution of mineral wool, glass wool, rock wool, and basalt wool at 37 C in water and in a Gamble’s solution modified by omission of the organic constituents. Most of the tests used a continuous-flow system in which the pH was 7.5–8. There was relatively little dissolution in distilled water in comparison to that produced by the modified Gamble’s solution. The surfaces developed a gel layer that, for the smaller diameters, extended throughout the fiber cross section. Thus, the fibers can become both smaller in outline and more plastic to deformation. Scholze and Conradt (1987) performed a comparative in vitro study of the chemical durability of SVFs in a simulated extracellular fluid under flow conditions. Seven vitreous, three refractory, and three natural fibers were involved. Samples of the leachate were analyzed, and the silicon concentrations were used to roughly classify the fibers according to their chemical durability in terms of glass network dissolution. A durability ranking of fiber materials was expressed in terms of a characteristic time required for the complete dissolution of single fibers of given diameter. SVFs exhibited relatively poor durability (with network dissolution velocities ranging from 3.5 to 0.2 nm per day for a glass wool and an E glass fiber, respectively), whereas natural fibers were very persistent against the attack of the biological fluid (e.g., less than 0.01 nm per day for crocidolite). Johnson et al. (1984) exposed rats to SVF aerosols at 10 mg/m3 for 7 h/day, 5 days/week for 1 year as compared to the single exposure of several hours duration used by Morgan and Holmes. The percentage of glass fibers with diameters less than 0.3 mm that were recovered from the lungs was consistently less than that in the original fiber suspension, and the reduction was more marked in the animals that were sacrificed following a period of recovery from the exposures than from those sacrificed at the end of the exposure. The degree of fiber etching increased with residence times in the lungs. Glass wool with and without resin was also etched, but to a lesser extent, and the etching of the rock wool fibers was considerably less. Bellmann et al. (1986) instilled reference suspensions of UICC crocidolite and chrysotile A, as well as suspensions of glass fibers in rat lungs and examined the residual fibers after 1 day and 1, 6, 12, and 15 months. Crocidolite fibers longer than 5 mm did not decrease in number for over 1 year. The number of chrysotile fibers >5 mm doubled, probably as a result of longitudinal splitting, while the number of glass fibers >5 mm was reduced by dissolution, with a half-time of 55 days. All fibers 20 mm) declined at a rate consistent with their exposure to a neutral pH environment. The diameters of shorter fibers declined much more slowly, consistent with exposure to the more acidic environment found in the phagolysosomes of alveolar macrophages. In the peritoneal cavity, all fibers, regardless of length, dissolved at the same rate as short fibers in the lung. The effect of dose on the distribution of fibers in the peritoneal cavity was investigated using experimental glass fibers and a powder made from ground fibers. At doses up to 1.5 mg, both fibers and powder were taken up by the peritoneal organs in proportion to their surface area, and uptake was complete 1–2 days after injection. At higher doses, the majority of the material in excess of 1.5 mg formed clumps of fibers (nodules) that were either free in the peritoneal cavity or loosely bound to peritoneal organs. These nodules displayed classic foreign body reactions, with an associated granulomatous inflammatory response. Collier et al. (1997) reported on the clearance of two stone wool fibers administered to rats by intratracheal instillation; one a conventional product (MMVF21) and the other an experimental, more soluble fiber (HTN). Unlike glass wool, stone wool is more soluble at the acid pH in macrophages than in the more neutral lung tissue. They found that MMVF21 had relatively slow clearance, with somewhat faster clearance for short fibers. The clearance of HTN was much faster. Eastes and Hadley (1995) administered suspensions of fibers to rats by intratracheal instillation, and the numbers, lengths, and diameters of fibers recovered from the lungs were measured by PCOM at intervals up to 1 year. Five different glass fibers had dissolution rates ranging from 2 to 600 ng/cm2/h measured in vitro in simulated lung fluid at pH 7.4. For fibers longer than 20 mm, the peak diameter decreased steadily with time after instillation, at the same rate measured for each fiber in vitro, until it approached zero. Measurements of the total number of fibers remaining in the rats’ lungs at times up to 1 year after instillation suggest that few of the administered fibers were being cleared by macrophage-mediated transport via the conducting airways. They concluded that glass fibers longer than 20 mm are removed from the lung by dissolution at the same rate measured in vitro. In a study of dissolution of inhaled fibers by Eastes and Hadley (1995), rats were exposed for 5 days to four types of airborne, respirable-sized SVF and to crocidolite fibers. The SVFs included two glass wools, and one each of rock and slag wool. After exposure, animals were sacrificed at intervals up to 18 months, and the numbers, lengths, and diameters of a representative sample of fibers in their lungs were measured. Long fibers (>20 mm) were eliminated from the rats’ lungs at a rate predicted from the dissolution rate measured in vitro. The long SVFs were nearly completely eliminated in several months, whereas the long crocidolite asbestos fibers remained in significant numbers at the end of the study. The number, length, and diameter distributions of fibers remaining in the rats’ lungs agreed well with a computer simulation of fiber clearance that assumed that the long fibers dissolved at the rate measured for each fiber in vitro, and that the short fibers of every type were removed at the same rate as short crocidolite asbestos. Thus, long SVFs were cleared by complete dissolution at the rate measured in vitro, and short fibers did not dissolve and were cleared by macrophage-mediated physical removal. In an inhalation study, using nine fiber types, Bernstein et al. (1996) exposed rats to an aerosol (mean diameter of 1 mm) at a concentration of 30 mg/m3, 6 h/day for 5 days with postexposure sacrifices at 1 h, 1 day, 5 days, 4 weeks, 13 weeks, and 26 weeks. At 1 h

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following the last exposure, the nine types of fibers were found to have lung burdens ranging from 7.4 to 33 106 fibers/lung with geometric mean diameters (GMD) of 0.40–0.54 mm, reflecting the different bi-variate distributions in the exposure aerosols. The fibers cleared from the lungs following exposure with weighted half-lives ranging from 11 to 54 days. The clearance was found to closely reflect the clearance of fibers in the 5–20 mm length range. An important difference in removal was seen between the long fiber (L > 20 mm) and shorter fiber (L between 5 and 20 mm and L < 5 mm) fractions, depending upon composition. For all glass wools and the stone wools, the longer fibers were removed faster than the shorter fibers. It was found that the time for complete fiber dissolution based on the acellular in vitro dissolution rate at pH 7.4 was highly correlated (r ¼ 0.97, p < 0.01) with the clearance halftimes of fibers >20 mm in length. No such correlations were found with any of the length fractions using the acellular in vitro dissolution rate at pH 4.5. Examination of the fiber length distribution and particles in the lung from 1 h through 5 days of exposure indicated that, especially for those fibers that form leached layers, fiber breakage may have occurred during this early period. These results demonstrate that, for fibers with high acellular solubility at pH 7.4, the clearance of long fibers is very rapid. Eastes and Hadley (1996) fitted much of the data cited above into a mathematical model of fiber carcinogenicity and fibrosis. Their model predicts the incidence of tumors and fibrosis in rats exposed to various types of rapidly dissolving fibers in an inhalation study or in an ip injection experiment. It takes into account the fiber diameter and the dissolution rate of fibers longer than 20 mm in the lung, and predicts the measured tumor and fibrosis incidence to within approximately the precision of the measurements. The underlying concept for the model is that a rapidly dissolving long fiber has the same response in an animal bioassay as a much smaller dose of a durable fiber. Long, durable fibers have special significance, since there is no effective mechanism by which these fibers may be removed. In particular, the hypothesis is that the effective dose of a dissolving long fiber scales as the residence time of that fiber in the extracellular fluid. The residence time of a fiber is estimated directly from the average fiber diameter, its density, and the fiber dissolution rate as measured in simulated lung fluid at neutral pH. The incidence of fibrosis in chronic inhalation tests, the observed lung tumor rates, and the incidence of mesothelioma in the ip model, were all well predicted by the model, The model allows one to predict, for an inhalation or ip experiment, what residence time and dissolution rate are required for an acceptably small tumorigenic or fibrotic response to a given fiber dose. For an inhalation test in rats at the maximum tolerated dose (MTD), the model suggests that less than 10% incidence of fibrosis would be obtained at the maximum tolerated dose of 1 mm diameter fibers if the dissolution rate were greater than 80 ng/cm2/h. The dissolution rate that would give no detectable lung tumors in such an inhalation test in rats is much smaller. Thus, a fiber with a dissolution rate of 100 ng/cm2/h has an insignificant chance of producing either fibrosis or tumors by inhalation in rats, even at the maximum tolerated dose. This model provides manufacturers of SVFs with design criteria for fibrous products that minimize, if not eliminate, their potential for producing adverse health effects. Support for the use of biopersistence data for the prediction of fibrosis and tumor responses in rats from both ip injection studies and chronic inhalation studies for fibers >5 and >20 mm in length was provided by Bernstein et al. (2001a, 2001b). For the inhalation studies they used collagen deposition at the broncho-alveolar junction as a predictor of interstitial fibrosis on the basis that it has been shown to be associated with tumor response in previous studies.

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12.5 PROPERTIES OF FIBERS RELEVANT TO DISEASE Fiber dimensions, chemical composition, and surface properties are important factors in biological reactivity of mineral fibers. As discussed previously, fiber length influences the deposition, clearance, and translocation of fibers in the lungs. Fiber length also determines clearance from the pleural and peritoneal spaces—fibers longer than 8 mm are trapped at the mesothelial lining because the opening of lymphatic channels draining these spaces are 8– 12 mm in diameter (Moalli et al., 1987). This provides an anatomic basis for the Stanton hypothesis that long fibers, regardless of their chemical composition, are more effective in producing mesotheliomas than shorter fibers, after direct intrapleural or intraperitoneal injection into rodents (Stanton et al., 1977). Cations within the crystal lattice may affect the toxicity of asbestos fibers. Mg2þ ions on the surface of chrysotile asbestos are important in cytotoxicity and carcinogenicity; acidleached fibers are less active than native fibers (Monchaux et al., 1981). The Fe2þ and Fe3þ content of amphibole fibers may be important because these cations can catalyze the Fenton or Haber–Weiss reactions, generating highly toxic and potentially mutagenic ROS (Weitzman and Graceffa, 1984; Zalma et al., 1987). Asbestos fibers generate reactive oxygen and nitrogen species (ROS/RNS), causing oxidation and/or nitrosylation of proteins and DNA. The ionic state of iron within asbestos fibers influences the oxidant-inducing potential and its influence on macromolecules, signal conduction pathways, inflammation and proliferation (Shukla et al., 2003).

12.6 FIBER-RELATED DISEASES/PROCESSES Macrophages are the initial target cells of inhaled particles that are deposited in the terminal airways and alveolar spaces. Phagocytosis of mineral fibers by macrophages leads to generation of ROS, and release of lysosomal enzymes, arachidonic acid metabolites, neutral proteases, chemotactic factors, and growth factors (Adamson, 1997). The interactions between mediators released from macrophages and other inflammatory cells and the target cell populations can initiate a sequence of events culminating in: fibrosis of the lungs and pleura; bronchogenic carcinoma; and malignant mesothelioma. However, for shorter exposures (2 weeks) to chrysotile, the early fibrotic lesions in rats are gradually resolved over the course of the following year (Pinkerton et al., 1997). Diffuse, bilateral interstitial fibrosis of the lungs characterizes asbestosis, a disease that usually develops in humans after prolonged exposure to high doses of asbestos fibers. Progressive scarring of the alveolar walls due to increased proliferation of fibroblasts and deposition of collagen produces radiographic evidence of disease, and interferes with gas exchange, leading to disability and premature death. The sequence of events that lead to the development of asbestosis includes the following: 1. Asbestos fibers are phagocytized by alveolar and/or interstitial macrophages. 2. Release of ROS from alveolar macrophages causes acute injury to the alveolar epithelial lining. The importance of hydrogen peroxide in the pathogenesis of asbestosis was demonstrated by the protection against asbestos-induced pulmonary fibrosis provided by catalast conjugated to polyethylene glycol (Mossman et al., 1990a, 1990b).

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3. Phagocytosis of asbestos fibers by alveolar or interstitial macrophages also triggers increased synthesis and release of growth factors for fibroblasts. Growth factors are released from macrophages exposed to asbestos fibers in vitro or in vivo: a homologue of platelet-derived growth factor (PDGF) (Bauman et al., 1990) and transforming growth factor-b (TGF-b) (Kane and McDonald, 1993). These growth factors cause chemotaxis of inflammatory cells and fibroblasts, stimulation of collagen synthesis, and inhibition of collagen degradation. Another reaction to asbestos exposure is the development of acellular fibrous scars, called pleural plaques, on the parietal pleural lining and diaphragm. Asbestos exposure may also lead to pleural effusions or diffuse fibrosis of the visceral pleura, but these reactions cause little disability. The most important reaction of the pleural and peritoneal linings to asbestos fibers is development of diffuse malignant mesothelioma, a rare neoplasm that is most closely associated with occupational or environmental exposure to amphibole forms of asbestos after a long latent period (up to 30–60 years). There is no increased incidence in cigarette smokers or in workers with asbestosis. This malignant neoplasm has a variable histology, ranging from epithelial to fibroblastic or mixed patterns. Malignant mesothelioma has usually spread diffusely when first diagnosed and responds poorly to radiation or chemotherapy (Craighead, 1987). The following sequence of events is hypothesized to lead to the development of diffuse malignant mesothelioma. 1. Phagocytosis of fibers that reach the pleura or peritoneal lining by macrophages. 2. Release of ROS, causes acute injury to the mesothelial cell monolayer lining the pleural or peritoneal spaces. This injury can be prevented by coating the administered fibers with the iron chelator, deferoxamine, with exogenous superoxide dismutase, or with catalase (Goodglick and Kane, 1990; Kane and McDonald, 1993). 3. Acute injury to the mesothelial lining, which is repaired by proliferation of adjacent, uninjured mesothelial cells. Growth factors released from macrophages, following phagocytosis of asbestos fibers, may modulate mesothelial cell regeneration (Kane and McDonald, 1993). 4. Direct interaction of asbestos fibers with the regenerating mesothelial cell population, which may cause chromosomal aberrations and aneuploidy. Additional DNA damage may be produced by reactive oxygen species, especially the hydroxyl radical produced by the iron-catalyzed Haber–Weiss reaction (Barrett et al., 1989; Floyd, 1990). 5. Repeated episodes of mesothelial cell injury and regeneration may lead to the emergence of a subpopulation of autonomously proliferating cells. 6. Neoplastic mesothelial cells may produce growth factors that promote growth of an invasive tumor. It is well established that workers exposed to asbestos fibers have an increased risk of developing bronchogenic carcinoma, and that workers who also smoke cigarettes have a greater risk. Cancers can arise from the epithelial lining of the large airways or terminal bronchioles. Bronchogenic carcinomas have a variety of histologic appearances: adenocarcinoma, squamous cell carcinoma (presumably arising in areas of squamous metaplasia of the respiratory epithelium), large cell carcinoma, and small cell (oat cell) carcinoma. These are the same histologic types of cancer associated with cigarette smoking in the absence of asbestos exposure (Mossman and Craighead, 1987).

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Lung cancer and mesothelioma are known to occur in people without radiographic evidence of lung fibrosis. deKlerk et al. (1997) demonstrated that the level of radiographic fibrosis conferred additional risk for lung cancer beyond that associated with level of exposure, but that asbestosis was not a prerequisite for asbestos-associated cancer. Bronchogenic carcinomas that develop in cigarette smokers show multiple alterations in proto-oncogenes and tumor-suppressor genes. It is unknown whether similar molecular changes are present in those malignant tumors that result from cigarette smoking in combination with asbestos exposure. Most of the experimental evidence suggests that asbestos fibers act as a cocarcinogen or tumor promoter in the respiratory lining, in conjunction with multiple components of cigarette smoke that may act as initiators. Some of these effects of asbestos fibers on the tracheobronchial epithelium may be mediated by ROS, since they are decreased by addition of various scavenging enzymes (e.g., superoxide dismutase, catalase) to these in vitro model systems.

12.7 REVIEW OF BIOLOGICAL EFFECTS OF SIZE-CLASSIFIED FIBERS IN ANIMALS AND HUMANS The pathological effects produced by fibers depend upon both the characteristics of the fibers and their persistence at sensitive sites. A number of carefully designed studies have been performed in which the size distributions of fiber suspensions have been well characterized as well as their persistence and/or effects. 12.7.1

In vivo Exposures by Instillation into Animal Lung Airways

King et al. (1946) instilled 100 mg of Rhodesian chrysotile into rabbit lungs at monthly intervals. One group received fibers microtomed to a length of 15 mm, and another group received fibers cut to 2.5 mm in length. At this huge dosage level, both groups showed foreign body reactions in the lungs. The long fiber produced a nodular reticulinosis, whereas the short fiber produced a diffuse interstitial reticulinosis. Wright and Kuschner (1977) used short and long asbestos and SVF in intratracheal instillation studies in guinea pigs. With suspensions containing an appreciable number of fibers longer than 10 mm, all of the materials produced lung fibrosis, although the yields varied with the materials used. However, with equal masses of short fibers of equivalent fiber diameters, none produced any fibrosis. The yields were lower for the long glass fibers than for the long asbestos, and this was attributed to their lesser durability within the lungs. 12.7.2

In vivo Exposures in Animals via Intraperitoneal Injection

For fibers injected ip (Davis, 1976; Pott et al., 1976; Wagner et al., 1976), or for fibers placed in a pledget against the lung pleura, a similar kind of fiber size and composition dependence was observed (Stanton and Wrench, 1972). The yield of mesotheliomas varied with fiber diameter and length, and with dose, with very little response when long, thin fibers were not included. Asbestos fibers were more effective than glass in these studies also. At a dose of 2 mg of chrysotile, crocidolite, or glass fiber, Pott et al. (1976) found only slight degrees of fibrosis, but tumor yields of from 16 to 38% in rats. When the chrysotile was milled to the extent that 99.8% of the fibers were shorter than 5 mm, the dose required to produce a comparable tumor yield (32%) was 50 times greater.

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Various hypotheses have been proposed to account for the pathological effects produced by asbestos. One was the contamination of the surface by trace metal and/or organic carcinogens. However, the studies of Stanton and Wrench (1972) found that surface contaminants played no role in mesothelioma yield, and concluded that the carcinogenicity of asbestos and fibrous glass was primarily related to the structural shape of these fibrous materials rather than their surface properties. Miller et al. (1999a) reviewed the collective outcomes of 9 rat intraperitoneal injection studies involving fibers of amosite, silicon carbide, four vitreous products (100/475, MMVF10, MMVF21, MMVF22), and three refractory ceramic products (RCF1, RCF2, RCF4) on mesothelioma. They reported a link between the numbers of injected fibers >20 mm in length, and to the biopersistence in the rat lung of fibers >5 mm in length. 12.7.3

In vivo Exposures of Animals by Inhalation

The relative potencies of the various mineral forms of asbestos after inhalation exposure are still not firmly established. Crocidolite is generally considered the most hazardous because of its association with significant numbers of human mesotheliomas. The relative toxicity of various asbestos minerals have been compared in a variety of experimental inhalation studies on small animals, but the results are in apparent conflict. Wagner (1963) reported more asbestosis with amosite than with chrysotile in guinea pigs, rats, and monkeys. However, in studies involving inhalation exposures of rats for 1 day to 2 years, Wagner et al. (1974), found that amosite and crocidolite were the least fibrogenic of five types of UICC asbestos, the others being Canadian chrysotile, Rhodesian (Zimbabwe) chrysotile, and anthophyllite (see Table 12.3). Holt et al. (1965) found no differences in the fibrogenic potential of chrysotile, crocidolite, amosite, and anthophyllite. Davis et al. (1978) used UICC chrysotile A, amosite, and crocidolite in 12-month rat exposures at respirable mass concentrations comparable to those used by Wagner et al. (1974) and found a similar pattern; that is, chrysotile was the most fibrogenic, and amosite and crocidolite the least. Heitt (1978) exposed guinea pigs by inhalation for 9 and 18 days and also found that chrysotile was more fibrogenic than amosite. Davis et al. (1986b) subsequently repeated the protocol with amosite with fiber length both shorter and longer than UICC amosite. The short amosite produced virtually no fibrosis, whereas the long amosite was more fibrogenic than chrysotile. The most fibrogenic asbestos was tremolite (see Table 12.3). Berman et al. (1995) analyzed the lung tumor and mesothelioma responses from 13 of the chronic inhalation studies in rats performed by Davis and colleagues at the Institute of Environmental Medicine in Edinburgh in relation to new measurements of the fiber distributions on archived chamber atmosphere sampling filters. The measure most highly correlated with tumor incidence was the concentration of fibers >20 mm in length. Miller et al. (1999b) reviewed the collective outcomes of 18 rat inhalation studies involving fibers of amosite, silicon carbide, four vitreous products (100/475, MMVF10, MMVF21, MMVF22), and three refractory ceramic products (RCF1, RCF2, RCF4). The primary influences on biological responses was the number of fibers 20 mm in length, and the dissolution rate of the fibers. Another important observation was that in vivo and in vitro cell responses did not significantly predict the risk of cancer following inhalation. McConnell et al. (1999) exposed hamsters for 78 weeks to amosite and two different fibrous glasses (MMVF10a and MMVF33) at 250–300 f/mL for fibers >5 mm in length, and to amosite at 125 and 25 f/mL as well. MMVF10a produced only mild inflammation, while

417

0.25

0.32 0.37

10

10 10

Amostiteshort Amositelong

Davis et al. (1986)

0.38 0.38 0.38 0.42 0.42

NR NR

10 10

10 10 5 10 2

NRcj NR NR

dm (mm)

10 10 10

Conc (mg/m3)

Tremolite

Amosite Crocidolite Crocidolite Chrysotile Chrysotile

UICC

Amosite Anthophvillite Crocidolite Chrysotile Canadian Rhodesian

UICC

Fiber Type

Resp.

30

1.7

28

16 12 12 30 30

NR NR

NR NR NR

>5 mm

10

0.1

7

2.7 4 4 16 16

NR NR

NR NR NR

>10 mm

Length (%)

Fiber Parameters

40

42

39

43 40 43 40 42

137 144

146 145 141

No. Rats

3 (7.5)

1 (2.4)

2 (5.1)

0 0 0 0 1 (2.4)

4 (2.9) 0

1 (0.7) 2 (1.4) 4 (2.8)

(4.7) (2.5) (4.7) (18) (14)

3 (7.5)

0

2 (5.1)

2 1 2 7 6

20 (15) 19 (13)

19 (13) 22 (15) 26 (18)

Adenoma

3 (7.5)

0

8 (21)

0 0 0 6 (15) 1 (2.4)

11 (8.0) 19 (13)

5 (3.4) 8 (5.5) 7 (5.0)

Adeno-CA

4 (10)

0

8 (21)

0 0 0 2 (5.0) 1 (2.4)

6 (4.4) 11 (7.6)

6 (4.1) 8 (5.5) 9 (6.4)

Squam-CA

Number (%) of Rats with Tumors Mesothelioma

Summary of Results of Mineral Fiber Inhalation Studies in Rats

Davis et al. (1985)

Davis et al. (1978)

Wagner et al. (1974)

Reference

TABLE 12.3

6.0 5.8

4.3 6.2 4.8

11.0

0.15

14.5

2.6 1.4 0.8 9.2 3.5

%bi

(continued)

Interstitial Scoreah

Fibrosis

418

Chrysotileshort Chrysotilelong

0.17 0.18

10

12

5

53 44

2

0.7

12 7

>10 mm

Length (%) >5 mm

0.30 0.22

dm (mm)

10

10 10

Conc (mg/m3)

Resp.

Fiber Parameters

40

40

28 28

No. Rats

From Lippmann (1988). a Relative scale where 1, nil; 2, minimal; 4, slight; 6, moderate; 8, severe at 24 months. b Percentage of tissue with fibrosis at 27–29 months. c Not reported.

Davis (1987)

UICC

Wagner et al. (1985)

Crocidolite Erionite

Fiber Type

(Continued)

Reference

TABLE 12.3

3 (7.5)

1 (2.5)

0 27 (96)

Mesothelioma

8 (20)

1 (2.5)

0 0

Adenoma

6 (15)

6 (15)

0 0

Adeno-CA

Number (%) of Rats with Tumors

5 (13)

0

1 (3.6) 0

Squam-CA

Interstitial Scorea

Fibrosis

12.6

2.4

%b

REVIEW OF BIOLOGICAL EFFECTS OF SIZE-CLASSIFIED FIBERS

419

MMVF33 produced more severe inflammation and mild interstitial and pleural fibrosis, as well as one mesothelioma. Amosite produced severe pulmonary fibrosis and many mesotheliomas (3.6%, 25.9%, and 19.5% at the low, medium and high doses). The effects were most closely related to the retained fibers >20 mm in length, and inversely paralleled the in vitro dissolution rates. Hesterberg and Hart (2000) reviewed the results of chronic inhalation studies in rats of 8 SFVs (RCF1, thin E-glass, 475-glass, 901-glass, CT-glass, slag wool, stone wool, rock wool), as well as amosite and crocidolite. The aerosols contained 100 f/mL >20 mm in length. Rats were exposed for 6 h/day, 5 days/week for 2 years and hamsters for 1.5 years. They also had biopersistence data for 5-day inhalations with retention measurements extending over 1 year. The more biopersistant fibers were fibrogenic (rock wool), or fibrogenic and carcinogenic (amosite, crocidolite, RCF1, 475-glass, and thin E-glass). Cullen et al. (2000) compared the pathogenicity of amosite to that of two special purpose glass microfibers with low dissolution rates (104E and 100/475) in a study in which rats were exposed for 7 h/d, 5 days/week for 12 months to 1000 f/mL longer than 5 mm. In terms of mesothelioma and lung cancers produced after the exposures and twelve months without further exposure, 104E and amosite fibers were considerably more potent than 100/475 fibers. They attributed the lower pathogenicity of 100/475 to the greater leaching of its component elements while in the lungs. 12.7.4

Fibers Retained in the Lungs of Exposed Humans

It is generally believed that amphibole fibers account for much of the mesothelioma incidence among exposed workers, even when they are predominantly exposed to chrysotile, since amphibole fibers are more biopersistent. Pooley (1976) examined postmortem lung tissue from 20 workers with asbestosis in the Canadian chrysotile mining industry and found that amphibole and other fibers were present in 16 cases. In seven of these, they were more numerous than chrysotile. In a later study of lung asbestos in chrysotile workers with mesothelioma, Churg et al. (1984) reported that the concentration ratio between cases and controls was 9.3 for tremolite but only 2.8 for chrysotile. In a Norwegian plant using 91.7% chrysotile, 3.1% amosite, 4.1% crocidolite and 1.1% anthophyllite, Gylseth et al. (1983) reported that the percentage of chrysotile in lung tissue ranged between 0 and 9%, while the corresponding numbers for the amphiboles were 76–99%. Case et al. (2000) examined the relationships between asbestos fiber type and length in the lungs of chrysotile textile workers as well as miners and millers in Quebec. Despite the lower cancer risk for the Quebec workers, the chrysotile, tremolite, total amphibole, and total count of fibers longer than 18 mm were all highest in the Quebec workers. They concluded that the textile workers experience should not be used to assess the cancer risks in other cohorts. In their review of the assessment of mineral fibers from human lung tissue, Davis et al. (1986a) attributed the high amphibole/chrysotile ratios to the dissolution of chrysotile within lung tissue, and the generally poor correlation between dust counts and mesothelioma as likely to be due to the differences among the various asbestos types in the fraction that reaches the pleural surface. Amosite fibers need to be longer to produce pulmonary fibrosis and pulmonary tumors in experimental animals than to produce mesotheliomas after injection (Davis et al., 1986b). Davis (1987) notes that both chrysotile and amphibole asbestos fibers inhaled by rats are plentiful in the most peripheral alveoli bordering on the pleura, but penetration of the external elastic lamina of the lung appears to be a rare event. On the contrary, erionite, a natural zeolite fiber, causes a very high incidence of mesotheliomas

420


in humans exposed to low environmental concentrations (Baris et al., 1987) and 100% incidence in rats exposed by inhalation (Wagner et al., 1985). Davis (1987) reported that the erionite used by Wagner et al. (1985) had a general appearance and fiber size distribution very close to that of UICC crocidolite, which produces a much smaller mesothelioma yield in rats exposed by inhalation. He attributed the difference to the enhanced ability of erionite to cross the pleural membrane.

12.8 CRITICAL FIBER PARAMETERS AFFECTING DISEASE PATHOGENESIS 12.8.1

Critical Fiber Dimensions for Asbestosis

Asbestosis has been caused by exposure to high concentrations of respirable fibers of all of the commercially exploited kinds of asbestos. Within the respirable fraction, the fibers often differ in diameter and length distributions and in retention times. The influence of these variables on fibrosis in the human lung was systematically explored and described by Timbrell et al. (1987) through analyses of both retained fibers and fibrosis in lung samples from exposed workers. He analyzed the fiber distributions in 0.5 g lung samples from several hundred workers by a technique known as magnetic alignment and light scattering (MALS) that he had previously developed (Timbrell, 1982). Optical microscopy was applied to measure the degree of fibrosis in paraffin sections prepared from adjacent samples of the same specimens. Wide intra- and intersubject variations were observed in fiber concentration and fibrosis score. The quantitative relationships between fibrosis and amphibole fibers in mineworkers’ postmortem lung specimens were determined for the main sources of amosite (Transvaal, South Africa), anthophyllite (Paakkila, Finland), and crocidolite (NW Cape and Transvaal, South Africa and Wittenoom, Australia). As illustrated in Fig. 12.3, the fibrosis-producing ability of the fibers was independent of amphibole type when normalized by the total surface area of long resident fibers per unit weight of lung tissue, presumably because the surface area determined the magnitude of the fiber–tissue interface. The wide range of the concentrations of retained fiber required to produce the same degree of fibrosis in the groups of mineworkers, when fiber quantity is expressed as number or total mass stemmed from the large differences between the distributions in diameter and length of the airborne fibers. Although the main focus of the Timbrell et al. (1987) study was on amphibole asbestos, they also reported results for three Wittenoom workers whose dominant exposure was to chrysotile asbestos. For these workers, chrysotile produced a similar degree of fibrosis to Wittenoom crocidolite for equal fiber mass concentrations in the lungs. Long residence in the tissue had almost completely dispersed the chrysotile fibers into fibrils, to give them a ratio of total surface area to mass resembling that of the particularly fine Wittenoom crocidolite fibers. The result indicates that the fibrogenicity of the retained chrysotile per unit of surface area within the lungs was similar to that of the amphiboles. Timbrell et al. (1987) also reported that amphibole mineworkers with a given fiber mass concentration in their lungs showed much higher degrees of fibrosis than gold miners with roughly the same mass concentration of retained quartz grains. The amphibole and quartz produced about the same fibrogenicity per unit of surface area, but the smaller diameters and higher area/mass ratios of the amphibole fibers endowed them with the greater surface area and thereby the superior fibrosis-producing capability.

CRITICAL FIBER PARAMETERS AFFECTING DISEASE PATHOGENESIS

421

FIGURE 12.3 Relationships between lung fibrosis scale and relative concentrations of fibers per unit weight of dry lung tissues. The lines connect data points from the same subject. The relative fiber surface area normalizes the data better than either the relative fiber number concentration of the fiber mass concentration. (Source: Lippmann, 1988.)

Knowledge of the interrelationships between retained fibers and fibrosis is critical in understanding the pathogenesis of the disease but is inadequate, by itself, in evaluating exposures to airborne fibers. This was recognized by Timbrell (1983, 1984), who developed a mathematical model relating fiber deposition and retention to analyses of lung samples. Specifically, he used samples from a woman at Paakkila who worked at a job that gave her exposure to an amphibole (anthophyllite) at high concentrations of fibers with a range of diameters and lengths sufficiently wide to encompass the size limits of respirable fibers. Her lungs contained 1.3 mg fiber/g of dry tissue, and she had asbestosis. One lung sample contained a fiber distribution matching the expected deposition. Timbrell speculated that severe fibrosis in the tissue in this sample had blocked the macrophage-mediated clearance. Another sample from the same lung yielded a retention pattern more closely matching those found in other Paakkila workers, with small fiber burdens and virtually no short fibers. He assumed that the latter represents long-term retention in the normal lung.

422


From the differences in retention, Timbrell developed a model for the retention of fibers as a function of length and diameter. Fiber retention rises rapidly with fiber lengths between 2 and 5 mm and peaks at 10 mm. Fiber retention also rises rapidly with fiber diameters between 0.15 and 0.3 mm, peaks at 0.5 mm, and drops rapidly between 0.8 and 2 mm. The utility of the model was demonstrated by applying it to predict the lung retention of Cape crocidolite and Transvaal amosite workers on the basis of the measured length and diameter distributions of airborne fibers. The predicted lung distributions did, in fact, closely match those measured in lung samples from a Cape worker (Timbrell, 1984) and, as shown in Fig. 12.4, from a Transvaal worker (Timbrell, 1983). Thus, fibrosis is most closely related to the surface area of fibers with diameters between 0.15 and 2 mm and lengths greater than 2 mm. The work of King et al. (1946) showing that chrysotile with length of 2.5 mm produced interstitial fibrosis in rabbits following multiple intratracheal instillations is consistent with the retention shown in Fig. 12.3 and a critical fiber length of 2 mm. Churg et al. (2000) examined putative biological mechanisms involved in fibrogenesis and conclude that: (1) fiber length, biopersistance, and dose it remains uncertain whether alveolar macrophages are central to fibrosis or whether fibers penetrating tissue are the real effector agents; (2) short fibers, readily degraded fibers, and small numbers of any fibers are nonfibrogenic; and (3) the ability of macrophages to clear fibers is probably crucial to preventing fibrosis.

12.8.2

Critical Fiber Dimensions for Mesothelioma

A National Research Council study (NRC, 1984) summarized mortality data for mesothelioma and lung cancer in asbestos-exposed occupational cohorts. In 20 studies in which there was an excess in respiratory cancer and/or mesothelioma, the percentage of the excess that was mesothelioma varied from 0 to 100%, with a mean (SD) of 38 29%. A study with 0% was that of Meurman et al. (1974, 1979), who reported 44 observed lung cancers (versus 22 expected) in a population of 1045 workers exposed to anthophyllite in Finland. Anthophyllite is an amphibole with larger fiber diameters than other forms of asbestos. By contrast, in several occupational cohorts the mesotheliomas accounted for more that 70% of the total. These included: (1) the study of Newhouse et al. (1982) of 7474 British workers exposed to mixed asbestos, among whom there were eight mesotheliomas and only three more than the 140 expected lung cancers; (2) the study of Rossiter and Coles (1980) of 6076 British shipyard workers exposed to mixed asbestos, among whom there were 31 mesotheliomas and 13 fewer lung cancers than the expected number of 101; (3) the study of Jones et al. (1980) of 578 British female workers exposed to crocidolite, among whom there were 17 mesotheliomas and six lung cancers more than the six expected; and (4) the study of Newhouse et al. (1982) of 3708 British female workers exposed to mixed asbestos, among whom there were two mesotheliomas and five fewer lung cancers that the 11 expected. Timbrell (1983), Timbrell et al. (1987), and Harington (1981) have noted that animal inoculation experiments have been interpreted as suggesting a fairly high value of diameter, for example, 1.5 mm (Stanton et al., 1977), 1 mm (Pott et al., 1976; WHO, 1986) and 0.25 mm (Wagner and Pooley, 1986), below which a fibrous material, so long as it is durable in lung fluids, can produce mesothelioma. In their view, these diameter limits are too high for human fiber-induced mesothelioma. If fibers with diameters >0.5 mm produced mesothelioma, then Paakkila, where the dust clouds contained on the order of 50 fibers/mL (PCOM) and a high proportion of fibers in the 0.5–3 mm diameter range, should have produced many


423

FIGURE 12.4 Distributions of fiber lengths and diameters of amosite asbestos in the lungs of a Transvaal worker. The predicted distribution at the left is based on the lengths and diameters of the airborne fibers and on the lung retention as a function of length and diameter. This corresponds closely to the distribution in the right panel, which was measured in samples from the worker’s lung. (Source: Lippmann, 1988.)

mesotheliomas, as well as excesses in fibrosis and lung cancer. As noted earlier, an average of 38% of the excess lung cancer plus mesothelioma in working populations exposed to asbestos was expressed as mesothelima. Despite the very high exposures of the Paakkila population, few mesotheliomas were observed. Timbrell (1983)Timbrell’s (1983) examination of the size distributions and mesothelioma incidence at Paakkila and other asbestos mines worldwide led him to conclude that a good correlation was obtained if the threshold diameter was reduced to 0.1 mm. The mesotheliomas that Paakkila fiber has produced in animals were, most likely, caused by the use of

424


excessive doses, 10,000 times that observed in man. Paakkila asbestos contains only 1% of fibers with diameters below 0.1 mm, but with such a large dose this represents an enormous absolute number. Harington (1981) noted that the data for the northwest Cape in South Africa, where numerous mesotheliomas have been reported, and for the northeastern Transvaal, where mesotheliomas are rare, are consistent with a low fiber-diameter limit. In the northwest Cape, about 60% of the fibers have diameters 5 mm, 2.5% > 10 mm) produced 30 mesotheliomas among 32 rats, and long amosite (30% > 5 mm, 10% > 10 mm) produced 20 mesotheliomas among 21 rats. Thus, fibers shorter than 5 mm appear to be ineffective, and an appreciable fraction longer than 10 mm appears to be unnecessary. 12.8.3

Critical Fiber Dimension for Lung Cancer

Excess incidence of lung cancer has been reported for workers exposed to amphiboles (amosite, anthophyllite and crocidolite), to chrysotile, and to mixtures of these fibers (NRC, 1984), but these studies have been uninformative with respect to the fiber parameters affecting the incidence. The series of rat inhalation studies performed by Davis et al. (1978), which have also produced lung cancers, have provided the most relevant evidence on the importance of fiber length on carcinogenicity in the lung. The Wagner et al. (1974) study found that the yield of squamous cell carcinoma and adenocarcinoma was greatest with Rhodesian chrysotile, with decreasing yields for Canadian chrysotile, crocidolite, anthopyllite, and amosite, respectively. As shown in Table 12.3, Davis et al. (1978) reported two squamous cell carcinomas, six adenocarcinomas and seven adenomas in 40 rats exposed to 10 mg/m3 of respirable chrysotile. In 42 rats exposed to


TABLE 12.4

425

Mesotheliomas Produced by Asbestos in Chronic Rat Inhalation Studies

Type of Asbestos Zimbabwe chrysotile Wagner et al. (1974) Davis et al. (1978) Davis et al. (1980)

Source Tumors/Animals 10 mg/m3 10 mg/m3 2 mg/m3 10 mg/m3 (1 day/week)

UICC UICC UICC UICC

Overall Quebec chrysotile Wagner et al. (1974) Davis et al. (1988) Hesterberg et al. (1993)

0/44 0/40 1/42 0/43 1/169 (0.6%)

10 mg/m3 10 mg/m3 10 mg/m3 10 mg/m3

UICC Short Long NIEHS

Overall Davis–Wagner Subset

4/44 1/40 3/40 1/69 9/193 (4.7%) 8/124 (6.5%)

Amphiboles Wagner et al. (1974) Crocidolite Amosite Anthophyllite

10 mg/m3 10 mg/m3 10 mg/m3

UICC UICC UICC

2/44 0/46 2/46

Davis et al. (1978) Crocidolite Crocidolite Amosite

5 mg/m3 10 mg/m3 10 mg/m3

UICC UICC UICC

1/43 0/40 0/43

Davis et al. (1980) Amosite

50 mg/m3 (1 day/week)

UICC

0/44

Wagner et al. (1985) Crocidolite

10 mg/m3

UICC

1/24

Davis et al. (1985) Tremolite

10 mg/m3

Korea

2/39

10 mg/m3 10 mg/m3

Short Long

1/42 3/40

10 mg/m3

UICC

1/69

Davis et al. (1986) Amosite

McConnell (1994) Crocidolite Overall Davis–Wagner subset

13/520 (2.5%) 12/451 (2.7%)

426


2 mg/m3 of chrysotile, there were six adenomas, one adenocarcinoma and one squamous cell carcinoma. There were also adenomas in the groups exposed to amosite at 10 mg/m3 (two) and to crocidolite (one at 10 mg/m3, two at 5 mg/m3). Davis et al. (1978) attempted to examine the influence of fiber number concentration in relation to mass concentration in their inhalation studies. Their five exposure groups included three at the same respirable mass concentration of 10 mg/m3, one each with chrysotile, crocidolite, and amosite. Of these, the amosite produced the lowest number concentration of fibers >5 mm in length. This fiber count was then matched with crocidolite (5 mg/m3 respirable mass) and chrysotile (2 mg/m3 respirable mass). In attempting to explain the greater fibrogenic and carcinogenic responses in the chrysotile-exposed animals than the crocidolite- or amosite exposed groups, they suggested it might have resulted, at least in part, from the greater number of >20 mm long fibers in the chrysotile aerosol. The ratio of >20 to >5 mm long fibers in the chrysotile was 0.185 compared to 0.040 for crocidolite and 0.011 for amosite. The diameter distributions of all three types of asbestos were similar, with a median diameter of 0.4 mm. The importance of fiber length to toxicity was further demonstrated by Davis et al. (1986b) on the basis of inhalation studies with amosite aerosols that were both shorter and longer than the UICC amosite studied earlier with the same protocols. Both aerosols had median diameters between 0.3 and 0.4 mm. The short-fiber amosite (1.7% > 5 mm in length) produced no malignant cancers in 42 rats, whereas the long-fiber amosite (30% > 5 mm, 10% > 10 mm) produced three adenocarcinomas, four squamous carcinomas and one undifferentiated carcinoma in 40 rats. In terms of adenomas, the frequencies were 3/40, 2/43, 0/42, and 1/81 for the long, UICC, short, and control groups, respectively. Davis et al. (1985) also studied tremolite asbestos using the same protocols. Its length distribution was similar to those of the chrysotile in the 1978 study and the long amosite in the 1986 study (i.e., 28% > 5 mm, 7% > 10 mm), but its median diameter was lower, that is, 0.25 mm. It produced two adenomas, eight adenocarcinomas, and eight squamous carcinomas in 39 rats. Davis (1987) reported on a study comparing the carcinogenic effects of “long” and “short” chrysotile at 10 mg/m3. Unfortunately, the discrimination between “long” and “short” fibers was less successful than that achieved for amosite. PCOM fiber counts for the fibers >10 mm in length for the “long” and “short” chrysotile were 1930 and 330 f/mL, whereas for the amosite they were 1110 and 12 f/mL, respectively. Despite the much more rapid clearance of the chrysotile from the lungs, the tumor yields were higher. For the “long” fiber, there were 22 tumors for the chrysotile versus 13 for the amosite. For the “short” fiber, there were seven versus none. Davis (1987) concluded that fibers 5 mm, 11.6% > 10 mm, median diameter 0.30 mm) or Oregon erionite (44% > 5 mm, 7.4% > 10 mm, median diameter of 0.22 mm). The UICC crocidolite produced one squamous carcinoma in 28 rats (but no mesotheliomas), whereas the erionite produced no carcinomas in 28 rats but did produce 27 mesotheliomas. In summary, Table 12.3 shows that 10 mg/m3 of short amosite (0.1% > 10 mm), UICC amosite (2.5% > 10 mm), UICC crocidolite (3% > 10 mm), and Oregon erionite (7.4% > 10 mm) failed to produce malignant lung cancers, whereas 10 mg/m3 of UICC chrysotile, long amosite and tremolite (all with 10% >10 mm) all produced malignant lung tumors. Although there was no clear-cut influence of fiber diameter on tumor yield, the results suggest that carcinogenesis incidence increases with both fiber length and diameter. Since Timbrell (1983) has shown that fiber retention in the lungs peaks between 0.3 and


427

0.8 mm diameter, it is likely that the thinner fibers, which are more readily translocated to the pleura and peritoneium, play relatively little role in lung carcinogenesis. Therefore, it appears that the risk of lung cancer is associated with long fibers, especially those with diameters between 0.3 and 0.8 mm, and that substantial numbers of fibers >10 mm in length are needed. In my own review of the literature on the chronic rat inhalation studies with amosite, brucite, chrysotile, crocidolite, erionite, and tremolite (Lippmann, 1994), I found that, for lung cancer, the percentage of lung tumors (y) could be described by a relation of the form y ¼ a þ bf þ cf2, where f is the number concentration of fibers, and a, b, and c are fitted constants. The correlation coefficients for the fitted curves were 0.76 for >5 mm f/mL, 0.84 for >10 mm f/mL, and 0.85 for >20 mm f/mL, and seemed to be independent of fiber type. This supports the hypothesis that the critical length for lung cancer induction is in the 10–20 mm range. In terms of the critical sites within the lungs for lung cancer induction, it has been shown that brief inhalation exposures to chrysotile fiber produces highly concentrated fiber deposits on bifurcations of alveolar ducts, and that many of these fibers are phagocytosed by the underlying type II epithelial cells within a few hours. Churg (1994) has shown that both chrysotile and amphibole fibers retained in the lungs of former miners and millers do not clear much with the years since last exposure. Thus, lung tumors may be caused by that small fraction of the inhaled long fibers that are retained in the interstitium below small airway bifurcations, where clearance processes are ineffective. One reason that short fibers may be less damaging could be the fact that they can be fully ingested by macrophages (Beck et al., 1971), and can therefore be more rapidly cleared from the lung. The fibrogenic response to long fibers could result from the release of tissuedigesting enzymes from alveolar macrophages whose membranes are pierced by the fibers they are attempting to engulf (Allison, 1977). The fibers may also cause direct physical injury to the alveolar membrane. A positive association of asbestosis with lung tumors was demonstrated by Wagner et al. (1974). The induction of fibrosis impairs clearance of deposited fibers, increasing the persistence of fibers in the lung. The preceding implies that short fibers will have a low order of toxicity within the lung, comparable to that of nonfibrous silicate minerals. Within this concept, the critical fiber length would most likely be on the order of the diameter of an alveolar macrophage, that is, about 10–15 mm. This line of reasoning leads to the same conclusion reached on the basis of the incidence of lung cancer in rats exposed to fibrous aerosols, that is, that the hazard is related to the number of fibers longer than 10 mm deposited and retained in the lungs. The Timbrell (1983) model predicts alveolar retention of deposited fibers approaching 100% for 10 mm long fibers in the 0.3–0.8 mm diameter range. Airborne fibers longer than 100 mm may be much less hazardous than those in the 10–100 mm range because they do not penetrate deeply into the airways, as interception increases with fiber length. 12.8.4

Summary of Critical Fiber Parameters

The various hazards associated with the inhalation of mineral fibers, that is, asbestosis, mesothelioma, and lung cancer, are all associated with fibers with lengths that exceed critical values. However, it now appears that the critical length is different for each disease, that is, 2 mm for asbestosis, 5 mm for mesothelioma, and 10 mm for lung cancer. There are also different critical values of fiber diameter for the different diseases. For asbestosis and lung cancer, which are related to fibers retained in the lungs, only fibers with diameters >0.15 mm need to be considered. On the contrary, for mesothelioma, which is initiated by fibers that

428


migrate from the lungs to the pleura and peritoneum, the hazard has been related to fibers with smaller diameters. A study, by Dufresne et al. (1996) of the fibers in lungs of Quebec miners and millers with and without asbestosis is supportive of the critical influence of long fibers on fibrosis and cancer incidence. They found that mean concentrations were higher in cases than in the controls for chrysotile fibers 5–10 mm long in patients with asbestosis with or without lung cancer; for tremolite fibers 5–10 mm long in all patients; for crocidolite, talc, or anthophyllite fibers 5–10 mm long in patients with mesothelioma; for chrysotile and tremolite fibers 10 mm long in patients with asbestosis; and crocidolite, talc, or anthophyllite fibers 10 mm long in patients with mesothelioma. Cumulative smoking index (pack-years) was higher in the group with asbestosis and lung cancer but was not statistically different from the two other disease groups. Although all durable fibers of sufficient length can produce fibrosis and cancer, as documented in various animal studies, it appears that factors other than fiber size can influence the extent of the response. For example, inhaled erionite appears to be much more potent for mesothelioma in both humans and animals because of its greater ability to penetrate the pleural surface. On the contrary, the animal and human data appear to differ on the ability of inhaled chrysotile to induce mesothelioma. Animal data indicate that chrysotile produces as much or more mesothelioma than the amphiboles, whereas human data more often implicate amphiboles, even when the predominant exposures are to chrysotile. Examination of the fiber content of the lungs of asbestos workers and animals exposed by inhalation shows that chrysotile is cleared much more rapidly than the amphiboles. It breaks down within the lungs both by disaggregation into fibrils and by dissolution. The differences between the responses in animals and humans may be in relative persistence, that is, time of persistence of the long fibers in the lung relative to the time interval between exposure and the expression of the disease. In other words, the long fibers may be retained in the lung for a longer fraction of the lifespan in the rat. Although all durable fibers in the right size range can cause the asbestos-related diseases, they may have different potencies and need different concentration limits. The remainder of this discussion addresses the indices of exposure but not the concentration limits for the fibers that fall within the indices. The concentration limits warrant separate and further discussion. 12.8.5 Implication of Critical Fiber Parameters to Health Relevant Indices of Exposure Although the current occupational exposure index, based on PCOM for fibers with an aspect ratio >3 and a length >5 mm, was a reasonable choice, when it was made, for occupational exposures involving specific known fiber types, it is now apparent that it cannot provide a scientifically adequate index for any of the differing hazards resulting from exposures to chrysotile, the various amphibole fibers, other mineral fibers, and the various SVFs. Its most important inadequacies for occupational exposure evaluations in mining, milling, and manufacturing industry include: (1) thin fibers of health relevance, that is, those with widths 5 mm long fibers that are visualized. It has further limitations for occupational exposures in demolition, building renovation, asbestos remediation projects, emergency response to steam pipe explosions, building collapses, and so on, where most of the dust collected on the sampling filter is

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nonfibrous. The background dust makes the counting of fibers difficult, if not impossible, and the fibers that are seen have a variety of compositions and toxicities. Some of these limitations can be overcome by using analytical transmission electron microscopy and X-ray Diffraction Analysis for fiber counting and analysis. These measurement methods enable visualization of fiber widths down to 0.1 mm, and composition analysis of each individual fiber in the field of view. Unfortunately, such analyses are considerably more expensive than PCOM for routine analyses, and the standard laboratory protocols have not utilized the available capabilities to generate fiber size distribution data that recognize the rapidly increasing hazard with fiber length as length increases above 5 mm. There is an additional limitation of the TEM methodology endorsed by EPA when it is applied to most exposure assessments for evaluating carcinogenic risks to the general public, since its detection limit for asbestos fibers >5 mm in length is higher than the concentration considered acceptable by EPA. This problem was discussed by HEI-AR (1991) and is further summarized in the next section. The TEM methodology has traditionally been used in EPAendorsed protocols to enumerate the count of all fibers greater than 0.5 mm in length, which ends up in insufficient filter area scanning to get a statistically valid sample of the healthrelevant fibers longer than 5, 10, or 20 mm. EPA has been considering the adoption of new dimensional criteria for hazardous fibers based on a proposal by Berman and Crump (2001), and sponsored a Peer Consultation Workshop (ERG, 2003b) on the topic. The Berman and Crump index, based in large measure on modeling of the results of the chronic inhalation studies in rats for which detailed fiber length and width data were available (Berman et al., 1995). Zero risk is assigned to fibers 5 mm in length in such buildings of 0.0002 f/mL correspond to a lifetime risk of about 2 10 6. However, it should be noted that concentrations in buildings are seldom much higher than concentrations in the air outside the buildings, and therefore much of this small risk is related to the entry of outdoor fibers into the building with the ventilation air. In contrast to risk estimations based on human experience in occupational populations and logical extrapolations to background concentration levels, as in the model of Doll and Peto (1985) described above, Larson (2003) applied the EPA Proposed Guidelines for Carcinogen Risk Assessment (EPA, 1996) to a set of lung cancer mortality data to obtain a “safe” fiber concentration based on a default linear extrapolation to one excess death per one million people, as specified for carcinogenic hazardous air pollutants by the Clean Air Act. He found that the “safe” concentration was 1/1000 of ambient air background concentrations of asbestos fibers. Because the calculated “safe” level cannot be achieved, Larsen suggested that that his risk assessment techniques be used only for airborne carcinogens that have only anthropogenic sources. Perhaps because he is an EPA employee, he did not question the reliability of the EPA Proposed Guidelines, with their numerous conservative defaults (Lippmann, 2003). Our current inability to: (1) reliably measure the concentrations of health-relevant fibers at concentrations near background levels; and (2) reliably quantitate the risks, if any, of exposures at such levels, has often led to confusion, alarm, and misguided acts of risk avoidance. For example, removal of in-place asbestos insulation in schools and public buildings has often increased rather than decreased exposures to asbestos fibers for workers doing the remediation and for building occupants after the remediation. Another example was the inappropriate focus, following the collapse of the World Trade Center buildings, on asbestos fibers in air and residual dust as indices of health risk for rescue workers and volunteers, workers removing debris, and neighborhood residents, office workers, and

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service workers. The measured airborne fiber levels neither warranted that level of concern nor could be related to the health effects that were documented. 12.10.6

Synthetic Vitreous Fibers

Although there has been a significant advance in our knowledge about the deposition and elimination of SVFs and other fibers in recent years, as well as some new knowledge about exposure–response in controlled animal inhalation studies, some further concern about lung cancer among heavily exposed workers in industry, and some new insight into the critical fiber dimensions affecting disease pathogenesis, there are also many important questions which remain to be addressed. In some cases, the behavior and risks of airborne SVFs can be inferred from those of either compact particles or asbestos fibers. On the contrary, the validity of such inferences depends on some critical assumptions about the aerodynamic properties of the various fibers and about the responses of lung and mesothelial cells to such fibers. The differences may be critical, and more in vivo studies with SVFs should be performed in order to further clarify these issues. In the interim, we already know a great deal about the nature and extent of fiber toxicity and the factors that modify its expression. This knowledge provides a good basis for a fairly definitive risk assessment for SVFs. SVFs differ from asbestos fibers in several critical ways and tend to produce less lung deposition and more rapid elimination of those fibers that do deposit in the lungs. One difference is in diameter distribution. Except for glass microfiber, SVFs tend to have relatively small mass fractions in diameters small enough to penetrate through the upper respiratory tract. Asbestos, on the contrary, usually contains more “respirable” fiber. Furthermore, once deposited, the asbestos fibers may split into a larger number of long thin fibers within the lungs. SVFs rarely split but are more likely to break into shorter length segments. There are also differences in solubility among the fibers that affect their toxic potential, among both the asbestos types and the SVFs. Conventional glass fibers appear to dissolve much more rapidly than other SVFs and asbestos. Dissolution of glass fibers takes place both by surface attack and by leaching within the structure. The diameters are reduced and the structure is weakened, favoring break up into shorter segments. Since the smallest diameter fibers have the greatest surface-to-volume ratio, they dissolve most rapidly. Thus, the relatively small fraction of the airborne glass fibers having diameters small enough to penetrate into the lungs are the most rapidly dissolved within the lungs. The more durable and less soluble SVFs, that is, slag and rock wool, some specialty glasses, and ceramic fibers, require a higher degree of concern because of their longer retention within the lungs. In vitro studies and studies of dissolution in simulated lung fluids can be very useful in preliminary evaluations of the toxic potential of the various SVFs. On the contrary, the dissolution of SVFs in vivo depends on many additional factors that cannot readily be simulated in model systems. For example, the differences in solubility in vivo of long and short fibers noted by Morgan and Holmes (1984) were attributed to small difference in intracellular and extracellular pH. The mechanical stress on fibers in vivo may also contribute to their disintegration, and cannot readily be simulated in model systems. Thus, hazard evaluations of specialty product SVFs made for limited and specific applications should include detailed in vivo studies in which animals are exposed to appropriate sizes and concentrations of the fibers of interest.

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In the case of conventional fibrous glasses, we have sufficient information to conclude that the occupational health risks associated with the inhalation of fibers dispersed during their manufacture, installation, use, maintenance, and disposal are not measurable (Doll, 1987), and hence of an extremely low order. The health risk from casual and infrequent indoor air exposure of building occupants to relatively low concentrations of fibrous glass is therefore essentially nil. These judgments are based on a series of interacting factors, each of which individually leads to a far lower order of risk for conventional glass fibers than asbestos. Specifically, 1. Conventional glass fibers are less readily aerosolized than asbestos during comparable operations, as demonstrated by the much lower fiber counts measured at various industrial operations (Cherrie et al., 1986; Esmen, 1984). 2. A much smaller fraction of conventional glass fibers than asbestos fibers have small enough aerodynamic diameters to penetrate into lung airways (i.e., fibers with diameters below 3 mm) (Konzen, 1984). 3. The glass fibers that can penetrate into the lungs are much less durable within the lung than asbestos. They tend to break up into shorter segments, so that fewer fibers longer than the critical length limits are retained at critical sites. They also tend to dissolve, further reducing their retention (Bernstein et al., 1984). 4. The inherent toxicity of conventional glass fibers is much lower than that of asbestos fibers of similar dimension, as shown by studies in which fiber suspensions are applied directly to target tissues by intratracheal instillation (Wright and Kuschner, 1977) or application of a fiber mat to the lung pleura (Stanton and Wrench, 1972). In consideration of these factors, the risk for lung fibrosis is virtually nil unless there is continuous exposure at concentrations high enough to maintain a high level of lung burden for this relatively rapidly cleared type of particulate. The risk of lung cancer is also virtually nil unless there is continuous exposure to long fibers at high concentrations because of the relatively rapid breakup of long fibers into short fiber segments within the lungs. Finally, the risk of mesothelioma from inhaled conventional glass fibers is virtually nil under almost any circumstance. There are hardly any glass fibers thin enough to cause mesothelioma in the aerosols, and the very few that may be present would dissolve rapidly within the lungs.

12.11 KEY FACTORS AFFECTING FIBER DOSIMETRY AND TOXICITY: RECAPITULATION AND SYNTHESIS 12.11.1

Critical Fiber Properties Affecting Toxicity

Review of the in vitro studies clearly indicates that fiber length, diameter, and composition are critical determinants of biopersistence, cytotoxicity and cell transformation. A review of the in vivo animal studies, both by inhalation and injection, shows that fiber dimensions and composition are important factors affecting pathological measures such as fibrosis and cancer yields. Review of human exposure–response shows that the proportions of the different diseases caused by asbestos, that is, asbestosis, lung cancer, and mesothelioma, vary greatly among occupational cohorts and that the mesothelioma/lung cancer ratio tends to increase with decreasing fiber diameter for the durable amphibole forms of asbestos.

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12.11.2


Influence of Fiber Diameter

Fiber diameter affects airborne fiber penetration into and along the lung airways and thereby the initial deposition patterns. The aerodynamic diameters of mineral fibers are about three times their physical diameters (Timbrell, 1972; St€ ober et al., 1970). Thus, fibers with diameters larger than 3 mm will not penetrate in the lungs (Lippmann, 1990). Fibers with diameters 0.1 mm are less well retained in the lungs than larger fibers (Lippmann and Timbrell, 1990). Their large surface-to-volume ratio favors dissolution (Lippmann, 1990). Those sufficiently durable not to dissolve can readily penetrate the epithelial surface and be translocated to the lung interstitium and pleural surfaces. The fibers that remain in the lungs can cause fibrosis and lung cancer, and those durable fibers that are translocated to pleural surfaces can cause mesothelioma. Thus, for asbestosis and lung cancer, the upper fiber diameter limit is on the order of 3 mm. For mesothelioma, the upper fiber diameter limit is likely to be much less for two reasons. First, the thinner fibers penetrate to the gas-exchange region to a greater extent. Second, fibers thinner than 0.5 mm are translocated from the deposition sites to postnodal lymphatic channels more than the thicker fibers and thus reach any organ of the body (Oberd€ orster et al., 1988). 12.11.3

Influence of Fiber Length

Fiber length can also affect fiber penetration into and along the airways. As the length increases beyond 10 mm, the interception mechanism begins to significantly enhance deposition (Sussman et al., 1991a, 1991b). Thus, longer fibers have proportionately more airway deposition and less deposition in the gas-exchange region. Lung retention also increases markedly with increasing fiber length above 10 mm for biopersistent fibers, both on theoretical grounds and on the basis of analysis of residual lung dust in humans (Pooley and Wagner, 1988; Churg and Wiggs, 1987; Timbrell et al., 1987) and animals (Morgan, 1979). Furthermore, fibers shorter than about 6 mm in length can readily penetrate through tracheobronchial lymph nodes and be translocated to more distant organs (Oberd€ orster et al., 1988). Exact specification of the critical lengths for the different diseases remains difficult, since the experimental studies generally have had, of practical necessity, to use imperfectly classified fiber suspensions. Also, the experimental studies have used very large concentrations, and apportioning attribution of the cytotoxicity and pathology produced to the effects of fiber size versus dust overload phenomena is difficult. In other words, the results described in the in vivo section of this review would be consistent either with short fibers having a much smaller effect than long fibers or with their contributing to the growth of fibrotic lesions caused by the relatively few long fibers in the tail of the fiber length distribution. In any case, the fibers shorter than 5 mm have very much less toxicity; whereas cytotoxicity and disease increase with fiber length for fibers longer than 5 mm. 12.11.4

Influence of Fiber Composition

Comparative retention and toxicity studies with various kinds of asbestos and other fibrous minerals, ceramics and glasses indicate that properties other than fiber dimensions affect fiber retention and toxicity. Among these are solubility; specific surface area; surface

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electrical charges that may contribute to redox reactions generating active oxygen species; and so on. Thus dimensional characteristics alone, although important, are insufficient indicators of fiber toxicity. It is now time to revise the Stanton hypothesis, which acknowledges the critical importance of fiber length and diameter in biological responses, and recognize the importance of the other physical-chemical properties that impart biological potential to fibers. A major research need is a systemic exploration of the surface properties and factors affecting solubility of fibers in lung fluids and cells, so that due considerations can be given to fiber composition in hazard assessment. 12.11.5

Risk Assessment for Inorganic Fibers

For asbestos, there is general agreement that occupational exposures to all fibrous forms have caused asbestosis and contributed to excesses of lung cancers. For mesothelioma, it is accepted that inhalation of amphibole and erionite fibers in workers and the general population has been causal. It is also generally agreed that occupational exposure to chrysotile asbestos has been associated with cases of mesothelioma, but many believe that these cases were more likely due to the contamination of most commercial chrysotile with amphibole fibers, and if chrysotile fibers do cause mesothelioma, they are considerably less potent in that regard than amphibole fibers. More definitive conclusions will require studies having better descriptions of the fiber sizes and compositions. For SVFs, the International Agency for Research on Cancer (IARC, 2002) has, on the basis of their own review of the data on carcinogenic risk, concluded that: (1) for humans, there is inadequate evidence for the carcinogenicity of glass wool, continuous glass filament, rock (stone) wool/slag wool, and refractory ceramic fibers; (2) for experimental animals, there is inadequate evidence for the carcinogenicity of continuous glass filament, and for certain newly developed, less biopersistent fibers (X-607 and HT wools and A, C, F, and G), limited evidence for insulation glass wool, rock (stone) wool, slag wool, and more biopersistent fibers such as fiber H, and sufficient evidence for the carcinogenicity of special purpose glass fibers, including E-glass and “475” glass fibers, as well as for refractory ceramic fibers. For all inorganic fibers, as for other airborne toxicants, the dose makes the poison. However, for these fibrous toxicants, the physical form and properties can be as important, or more important, than the chemical form. The aerodynamic diameters of the fibers, and therefore their deposition patterns and efficiencies within the lung airways, is determined largely by fiber width. For fibers >10 mm in length, interception enhances airway deposition, but even more importantly, these longer fibers elicit cellular responses that shorter fibers do not, and they also are subject to different clearance pathways and rates. Another physical property, their solubility within lung fluids, then becomes a major determinant of their toxicity. While these special determinants of risk are being increasingly recognized, they are not yet reflected in Standards or Guidelines for exposure assessment, an essential tool in risk assessment. Thus, there is an urgent need for new and improved occupational and ambient air quality limits for specific fiber compositions that recognize fiber length and diameter as critical risk factors. Dissolution rate in vivo is the other main dimension in the risk equation. Fortunately, the SVF manufacturing industry in the U.S. and the European Community has recognized the critical importance of fiber biopersistence to risk, and has revised many product formulations so that they have higher dissolution constants.

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ACKNOWLEDGMENTS This research was performed as part of a Center Program supported by NIEHS (Grant ES, 00260). It includes extensive review material from earlier review papers, specifically Lippmann (1988, 1990, 1994 and HEI-AR, 1991) Health Effects Institute-Asbestos Research (1991).

12.13 ACRONYMS ACGIH: ACM: AM: BMRC: HEI-AR: ip MALS: MMMF: MMVF: MPPCF: MTD OSHA: PCOM: PEL: PMN: RCF ROS SIR SMR: SVF TEM: TLV: UICC:

American Conference of Governmental Industrial Hygienists Asbestos-containing material Alveolar macrophage British Medical Research Council Health Effects Institute-Asbestos Research intraperitoneal, a dose delivery technique for fibers Magnetic alignment and light scattering Manmade mineral fiber, an alternate name for SVF Manmade vitreous fiber, an alternate name for SVF Millions of particles per cubic foot Maximum tolerated dose Occupational Safety and Health Administration Phase-contrast optical method Permissible exposure limit Polymorphonuclear leukocytes Refractory ceramic fiber Reactive oxygen species Standardized incidence ratio Standardized mortality ratio Synthetic vitreous fiber Transmission electron microscopy Threshold limit value International Union Against Cancer (English translation of name of organization in French

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Saracci R, Simonato L, Acheson ED, erson A, Bertazzi PA, Claude J, Charnay N, Esteve JJ, FrentzelBeyme RR, Gardne MJ, Jensen OM, Massing R, Olsen JH, Teppo L, Westerholm P, Zocchetti C (1984) Mortality and incidence of cancer of workers in the man-made vitreous fibres producing industry: an international investigation in 13 European plants. Br. J. Ind. Med. 41:425–436. Sawyer RN (1989) Asbestos material inventory, control concepts and risk communications. In: Symposium on Health Aspects of Exposure to Asbestos in Buildings. Cambridge, MA: Harvard University Kennedy School of Government, pp. 155–169. Schlesinger RB, Lippmann M (1978) Selective particle deposition and bronchogenic carcinoma. Environ. Res. 15:424–431. Schlesinger RB, Concato J, Lippmann M (1983) Particle deposition during exhalation: a study in replicate casts of the human upper tracheobronchial tree. In: Marple VA, Liu BYH, editors. Aerosols in the Mining and Industrial Work Environments, Vol. 1. Ann Arbor: Ann Arbor Science Publishers, pp. 165–176. Scholze H, Conradt R (1987) An in vitro study of the chemical durability of siliceous fibers. Ann. Occup. Hyg. 31 (4B):683–692. Sebastien P, Bignon J, Baris YL, Awad L, Petit G (1984) Ferruginous bodies in sputum as an indication of exposure to airborne mineral fibers in the mesothelioma villages of Cappadocia. Arch. Environ. Health 39:18–23. Sebastien P, McDonald JC, McDonald AD, Case B, Harley B (1989) Respiratory cancer in chrysotile textile and mining industries: exposure inferences from lung analysis. Br. J. Ind. Med. 46:180–187. Seidman H, Selikoff IJ, Hammond EC (1979) Short term asbestos work exposure and long term observation. Ann. NY Acad. Sci. 330:61–90. Selikoff IJ, Nicholson WJ, Langer AM (1972) Asbestos air pollution. Arch. Environ. Health 25:1–13. Selikoff IJ, Hammond EC, Seidman H (1979) Mortality experience of insulation workers in the United States and Canada: 1943–1976. Ann. NY Acad. Sci. 330:91–116. Shannon H, Hayes M, Julian J, Muir D (1984) Mortality experience of glass fibre workers. In: Biological Effects of Man-Made Mineral Fibers. Copenhagen: World Health Organization, pp. 347–349. Shannon HS, Jamieson E, Julian JA, Muir DCF, Walsh C (1987) Mortality experience of Ontario glass fiber workers—extended follow-up Ann. Occup. Hyg. 31 (4B):657–662. Shukla A, Gulumian M, Hei TK, Kamp D, Rahman Q, Mossman BT (2003) Multiple roles of oxidants in the pathogenesis of asbestos-induced diseases. Free Radical Biol. Med. 34:1117– 1129. Simonato L, Fletcher AC, Cherrie J, Andersen A, Bertazzi PA, Charney N, Claude J, Dodgson J, Esteve J, Frentzel-Beyme R, Gardner MJ, Jensen O, Olsen J, Saracci R, Teppo L, Westerholm P, Winkelmann R, Winter PD, Zocchetti C (1986) Updating lung cancer mortality among a cohort of man-made mineral fibre production workers in seven European countries. Cancer Lett. 30:189– 200. Smith AH, Wright CC (1996) Chrysotile asbestos is the main cause of pleural mesothelioma. Am. J. Ind. Med. 30:252–266. Spengler JD, Ozkaynak H, McCarthy JF, Lee H (1989) Summary of Symposium on Health Aspects of Exposure to Asbestos in Buildings. In: Symposium on Health Aspects of Exposure to Asbestos in Buildings. Cambridge, MA: Harvard University Kennedy School of Government, pp. 1–26. Stanton MF, Wrench C (1972) Mechanisms of mesothelioma induction with asbestos and fibrous glass. J. Natl. Cancer Inst. 48:797–821. Stanton MF, Layard M, Tegeris A, Miller E, May M, Kent E (1977) Carcinogenicity of fibrous glass: pleural response in the rat in relation to fiber dimension. J. Natl. Cancer Inst. 58:587–603. Stayner LT, Dankovic DA, Lemon RA (1996) Occupational exposure to chrysotile asbestos and cancer risk: a review of the amphibole hypothesis. Am. J. Public Health 86:179–186.

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13 BENZENE Bernard D. Goldstein and Gisela Witz

Understanding and preventing the threat of benzene (C6H6) to human health is one of the most important environmental issues facing national and international regulatory authorities. Benzene causes human leukemia. Among the known human cancer-causing agents, benzene is the organic chemical of highest volume and broadest distribution. Further, as an integral component of our petrochemical era and a constituent of crude oil, benzene cannot simply be banned from use. Understanding the mechanisms by which benzene leads to adverse health effects is of crucial importance. Uncertainties about health effects must be balanced against the potential for substantial economic and societal costs in regulating benzene, as well as the potential adverse effects of benzene substitutes. Current standards for benzene in the United States include a maximum permissible level in drinking water of 5 ppb by weight. As for any carcinogen, the U.S. Environmental Protection Agency (EPA) has set a drinking water goal of 0 ppb. The U.S. Occupational Safety and Health Administration (OSHA) has set a workplace standard for benzene of 1.0 ppm benzene by volume as a time-weighted average for an 8 h working day. The U.S. National Institute for Occupational Safety and Health (NIOSH) has recommended a workplace air standard of 0.2 ppm. The control of benzene in ambient air by EPA is currently based on emission standards set for selected industrial sources. Under the 1990 Clean Air Act Amendments, maximum available control technology for all significant atmospheric benzene point sources is required, followed by a risk-based approach that has yet to be clearly defined. Many states have developed their own drinking water or atmospheric standards for benzene. International standards vary greatly. Benzene is the smallest and most stable aromatic compound. It is a clear colorless liquid with a classic aromatic odor. It is minimally soluble in water (820 mg/L at 22 C) and has an octanol/water partition coefficient of 1.56–2.15 (Leo et al., 1971), a blood/air partition coefficient of 7.8 (Sato and Nakajima, 1979), and a vapor pressure of 0.125 atm at 25 C (Thibodeaux, 1981). It is thus a hydrophobic solvent that readily evaporates at room temperature and rapidly partitions into lipid. Benzene reacts with hydroxyl radicals and


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participates in the photochemical process leading to the formation of ozone and other components of oxidant smog. Described below are a number of aspects of the toxicology and exposure pathways of benzene pertinent to understanding how the use of benzene in our modern society leads to the risk of adverse health effects. Pertinent review articles or documents include Goldstein (1977, 1989a, 1989b), Snyder (2002, 2004), Savitz and Andrews (1997), Smith (1996), Ross (2000, 2005), Lovern et al. (2001), Bird et al. (2005), Rana and Verma (2005), and Zhang et al. (2002). Both the U.S. Agency for Toxic Substances and Disease Registry (ATSDR) and EPA have draft revisions of their comprehensive benzene documents in the review stage, with publication expected in late 2006 or 2007.

13.1 BENZENE EXPOSURE Benzene is a ubiquitous agent. As a component of petroleum, it is widely distributed. Gasoline contains 1–2% benzene in the United States, and higher levels are reported elsewhere. Benzene is also an important starting agent for chemical synthesis. It is a valuable solvent, but its use in that regard has been decreasing, primarily because of health and safety concerns. Benzene exposure occurs in the workplace, in the general environment, and through the use of consumer products. Occupational exposures present the highest risks. Cigarette smoke contains relatively high levels of benzene. For nonsmokers, benzene sources in the home are usually the major component of exposure. Benzene is also present in some foods, but relatively little of the usual total daily body burden is likely to come from this source. The World Health Organization (WHO) estimates that total daily uptake from all sources ranges from 130 to 550 mg in nonsmokers (WHO, 1987). Much of what we know about the extent of individual benzene exposure in the general environment began with a series of pioneering studies by EPA’s Office of Research and Development. They developed miniaturized sampling and analytical techniques suitable for personal monitors, demographic sampling techniques to choose individuals representative of community exposure, and, most importantly, conceptual approaches allowing for integration of indoor and outdoor exposure data for individuals in conjunction with activity questionnaires (Wallace et al., 1985; Wallace, 1987). Perhaps the most startling information concerning benzene came from studies in northern New Jersey, near a major petrochemical refinery complex (Wallace et al., 1985; Wallace, 1987). Evaluation of 355 individuals failed to reveal any statistically significant impact of proximity to the refinery on individual benzene exposure. This does not mean there was no impact; outdoor monitors confirmed the human olfactory perception of higher outdoor levels of petrochemical vapors, including benzene, in proximity to the refinery complex. However, the most notable finding was the large variability in indoor benzene levels, a variability so great that it swamped the relatively small differences caused by geographical proximity to the refinery. For those with the higher individual levels of benzene exposure, that is, those at greatest risk, the indoor sources clearly predominated. Personal exposure is usually even higher than that predicted solely by indoor levels, reflecting activity patterns near benzene sources. In general, homes in northern communities have higher indoor benzene levels, reflecting the likelihood of both attached garages and restricted outdoor ventilation in the winter. The Total Exposure Assessment Methodology (TEAM) was also the basis for study of the risk of benzene exposure resulting from emissions from the Marine Oil Terminal in Valdez, Alaska. Despite exceptionally high emissions, personal monitoring

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coupled with extensive meteorological and tracer studies demonstrated that the Marine Oil Terminal contributed only minimally to the benzene cancer risk of the inhabitants living in Valdez, located 5 km away. Indoor sources predominated (Goldstein, 1994). The approaches pioneered by the TEAM study have continued to be adapted and improved, including use of breath analysis as a means of assessing body benzene burden (Yu and Weisel, 1996). Benzene may also be measured in blood, and benzene metabolites can be measured in urine (Ashley et al., 1994; Medeiros et al., 1997; Rappaport et al., 2005; Wallace et al., 2004). Many new techniques have been developed, such as protein adducts (Rappaport et al., 2005) and a variety of urinary and other markers (Qu et al., 2005). These techniques should be considered as state of the art when exploring individual human exposures. In particular, regulatory decisions made about a pollutant source should no longer depend solely on source-based mathematical modeling approaches to determine whether there is sufficient risk to warrant imposing control measures or closing the facility. In addition to background levels of benzene from industrial sources in the community, the general public may be exposed to benzene in a number of ways. Some of the major ones are as follows: 1. Cigarette smoking, including passive smoking: Benzene levels are elevated in areas with significant levels of cigarette smoke, and blood benzene levels have been shown to be higher in smokers than in nonsmokers (Wallace and Pellizari, 1986). One pack a day contributes about 600 mg benzene to the smoker (WHO, 1987). 2. Home use of solvents or gasoline: Many solvents contain benzene, in almost all cases at levels less than 0.1%. However, if allowed to evaporate freely, even at 0.1% (1000 ppm) solvents can be a measurable contributor to airborne benzene levels in the home. Gasoline, which contains 1–2% benzene, is often a major source of benzene at home, particularly for those who have gasoline-fueled machinery, such as automobiles or lawnmowers in attached garages. Other consumer products that are sources of benzene include household cleaning agents, art and other hobby supplies, and glues. 3. Leaky underground storage tanks: Water supplies have become contaminated with benzene as a result of leaks from underground gasoline storage tanks. Contamination of groundwater can lead to human exposure through three routes: ingestion, inhalation, and skin absorption. Inhalation and skin absorption can occur during such activities as showering with benzene-contaminated potable water (Weisel et al., 1996). Significant risk of benzene inhalation from groundwater contamination may occur even in situations in which a municipal water supply is unaffected. This can occur by off-gassing through basement walls and floors. In one instance on Long Island in New York, which has particularly porous sandy soils, a group of over 20 homes was sufficiently affected by a leaky underground storage tank from a nearby gasoline station that the gasoline levels in the basement reached potentially explosive concentrations and the families had to be evacuated. Eventually, the decision about when it would be safe to return the families to their homes was based on arguments concerning the leukemia risk of residual air benzene concentrations. As is common in these unfortunate situations, after a period of living in a motel the families preferred to move elsewhere rather than return to their homes. 4. Automotive sources: Automobile-related emissions remain a substantial source of community exposure to benzene. Evaporation of ambient gasoline from the fuel train within the car has been largely, but still incompletely, controlled. However, release of

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benzene and other gasoline vapors during refueling remains a significant source of community benzene exposure, as well as exposure to the individual doing the refueling. Benzene is also emitted during the combustion of gasoline. The catalytic converter installed on automobile exhausts is an effective means of reducing benzene emissions. Inhalation of benzene while in an automobile can be a significant portion of total daily benzene burden (Dor et al., 1995; Lawyrk and Weisel, 1996).

13.2 UPTAKE Absorption of benzene occurs through inhalation, ingestion, and across the skin. Except for unusual circumstances, inhalation of benzene is the major route of absorption. Benzene is readily absorbed in the lung, directly entering the bloodstream, where it is distributed to the tissues. Benzene within the blood is in direct equilibrium with the benzene in expired air. Thus, measurement of end alveolar breath benzene concentration is a good indicator of body benzene concentration. Approximately 50% of benzene taken up into the body by any route is eventually exhaled, the extent being dependent on benzene dose and the rate of metabolism and respiratory mechanics (i.e., assisted ventilation is effective in removing benzene from the body for treating benzene-induced acute central nervous system toxicity). Ingested benzene is also assumed to be fully absorbed. The skin is a more effective barrier. Studies of the absorption rate of liquid benzene across the skin have demonstrated significant uptake both in vitro and in vivo, with time in contact with the skin being a major factor (Franz, 1984; Loden, 1986). The extent of worker transdermal exposure to benzene is being explored using the new technique of charcoal cloth pads (Van Wendel de Joode et al., 2005). There is no evidence of transdermal absorption of benzene vapor. Much more needs to be done to understand the rate of skin absorption of benzene in liquid mixtures, such as gasoline and commercial solvents. Furthermore, it is at least theoretically possible that blends of gasoline with oxygenated fuels such as methyl tert-butyl ether or ethanol will lead to a more rapid rate of benzene absorption across the skin.

13.3 METABOLISM AND DISPOSITION Benzene is relatively inert to chemical additions, eliminations, oxidations, and reductions because it lacks substituents that can be altered and/or that confer chemical reactivity to the aromatic ring. Benzene is a nonpolar organic compound that partitions into fatty tissues. Numerous studies (Andrews et al., 1977; Sammett et al., 1979; Bolcsak and Nerland, 1983) indicate that benzene requires metabolism to reactive intermediates in order to be toxic. This discussion briefly reviews the essentials of benzene metabolism and disposition. 13.3.1

In Vivo Metabolism

The major pathways of benzene metabolism are shown in Fig. 13.1. The earliest studies of benzene metabolism in vivo reported the formation of phenol (Schultzen and Naunyn, 1867), catechol, and hydroquinone (Nencki and Giacosa, 1880). Porteous and Williams (1953) found that phenol, catechol, p-benzoquinone, and hydroquinone are excreted as ethereal sulfates in the urine of rabbits dosed orally with benzene. Earlier, Jaffe (1909) and other

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463

FIGURE 13.1 Structure of benzene and adducts investigated as potential biomarkers of benzene exposure. From Medeiros et al. (1997) and Taylor and Francis (1997), used with permission.

investigators (Drummond and Finar, 1938) reported the urinary excretion of muconic acid, a ring-opened six-carbon diene dicarboxylic acid, in rabbits. In l953, Parke and Williams, using [14 C]benzene, confirmed and extended the early studies on the metabolism of benzene to ring-hydroxylated metabolites and to trans,trans-muconic acid. In rabbits administered 0.3–0.5 mL/kg [14 C]benzene by gavage, the major metabolite formed was phenol. Catechol and hydroquinone were also detected. These, along with phenol, were eliminated in the urine, mainly as the ethereal sulfate or glucuronic acid conjugates. trans,trans-Muconic acid (muconic acid) was also detected in the urine. Labeled carbon dioxide, indicating benzene ring opening, and phenylmercapturic acid were also detected. Metabolic fate studies indicated that 43% of the administered benzene dose was expired unmetabolized, 1.5% was exhaled as CO2, 35% was recovered as urinary metabolites, and 5–10% was present in feces and body tissues. Urinary metabolites consisted of 23% phenol, 4.8% hydroquinone, 2.2% catechol, and 1–2% trans,trans-muconic acid. A second ring-opened metabolite, 6-hydroxy-trans,trans-2,4-hexadienoic acid (6-hydroxyhexadienoic acid, HHA), was recently identified by Kline et al. (1993) as a urinary metabolite of benzene in mice. In studies on benzene metabolism in the isolated perfused mouse liver, Hedli et al. (1997) found a significant difference between single-pass metabolism in the orthograde (normal) and that in the retrograde (reversed) direction. Although the amount of phenol plus its conjugates produced was the same regardless of the direction of perfusion, the amount of free phenol formed expressed as a percentage of total phenolic metabolites was twice as great following normal perfusion compared with reversed perfusion, indicating regional differences in the

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location of cytochrome P450 and the conjugation enzymes. Phenol conjugates and small amounts of free and conjugated hydroquinone, but no free phenol, were detected after recirculation of products formed during single-pass orthograde perfusion. These results could, in part, explain why administration of phenol does not lead to bone marrow depression. These major pathways of benzene metabolism, originally determined in rabbits, were subsequently established in rats and mice (Sabourin et al., 1988a; for reviews see Snyder, 1987; Snyder et al., 1993). In general, in vivo metabolism of benzene results in the formation of ring-closed metabolites and the ring-opened metabolites muconic acid and 6-hydroxyhexadienoic acid. The latter two compounds can be formed by metabolism of trans,trans-muconaldehyde (Witz et al., 1990; Goon et al., 1992; Zhang et al., 1993), a reactive ring-opened microsomal metabolite of benzene (Latriano et al., 1986). The metabolism to phenol and subsequent formation of phenyl glucuronide or sulfate is a detoxication pathway, as is conjugation with glutathione and subsequent formation of the prephenylmercapturic acid [S-(1,2-dihydro-2-hydroxyphenyl)-N-acetyl cysteine]. Quinol thioethers were recently reported to be present in bone marrow of mice and rats administered 11.2 mmol/kg benzene twice daily for 2 days (Bratton et al., 1997). The quinol thioethers identified consisted of 2-(glutathione-S-yl)hydroquinone [2-(GSyl)HQ], 2-(cystein-S-ylglycinyl)hydroquinone [2-(Cys-Gly)HQ], 2-(cystein-S-yl)hydroquinone [2-(Cys)HQ], and 2-(N-acetyl-cysteine-S-yl)hydroquinone [2-(NAC)HQ]. The metabolite 2-(GSyl)HQ is most likely derived by the reaction of glutathione with p-benzoquinone, a product formed by the oxidation of hydroquinone. Metabolism via the mercapturic acid pathway was demonstrated to lead to the quinol thioethers derived from 2-(GSyl)HQ. Metabolism of phenol to hydroquinone and that of benzene to muconic acid are two pathways that are currently thought to lead to the formation of toxic benzene metabolites. Hydroquinone and p-benzoquinone, easily formed through the oxidation of hydroquinone, are reactive metabolites that have been suggested to play a role in benzene toxicity (Schwartz et al., 1985; Irons, 1985; Sawahata et al., 1985; see Ross, 2005 for a recent review). Muconaldehyde, a putative precursor of urinary muconic acid, is hematotoxic in mice (Witz et al., 1985) and may be responsible, in part, for benzene toxicity. p-Benzoquinone– glutathione adduct formation leading to quinol thioethers could potentially represent yet another pathway resulting in the formation of toxic benzene metabolites. Several quinol thioether metabolites of hydroquinone have been shown to inhibit erythropoiesis in rats (Bratton et al., 1997). Potential mechanisms for their bone marrow toxicity could involve acylation of critical cellular molecules, as well as oxidative damage by reactive oxygen species generated via redox cycling (Bratton et al., 1997; Rao, 1996; Ross, 2000; Smith, 1996). 13.3.2

Mechanism(s) of Metabolite Formation

The liver and, to a lesser extent, the bone marrow and more recently the lung are the organ systems examined for benzene metabolism. Studies by Sammett et al. (1979) originally demonstrated that partial hepatectomy inhibits benzene hematotoxicity. It also decreases benzene metabolism by 70%. Coadministration of toluene, a competitive inhibitor of benzene metabolism, also reduces benzene hematotoxicity and decreases the amount of benzene metabolites excreted in the urine and found in the bone marrow, blood, liver, spleen, and fat tissue (Andrews et al., 1977). Based mainly on these studies, benzene toxicity is thought to be mediated by reactive metabolites that are formed via pathways including


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metabolism of benzene by the liver. This hypothesis does not preclude hepatic metabolism of benzene to intermediates that travel from the liver to the bone marrow, where they could be further metabolically activated to the ultimate myelotoxic species. Many in vitro studies utilizing cellular fractions or reconstituted purified enzyme systems have been carried out in order to elucidate the mechanisms of benzene metabolite formation. Benzene is initially metabolized by a cytochrome P450-dependent monooxygenase to phenol. The formation of phenol is believed to involve the intermediate formation of benzene oxide, followed by rearrangement to phenol (Jerina and Daly, 1974), or acid-catalyzed opening of the epoxide ring, followed by aromatization via loss of a proton. A direct insertion of oxygen for aromatic hydroxylation may also account for the formation of appreciable amounts of phenol (Hanzlik et al., 1984). Urinary S-phenylmercapturic acid and N7phenylguanine are adducts presumably derived from the reaction of benzene oxide with glutathione and DNA, respectively. Identification of these adducts (Parke and Williams, 1953; Norpoth et al., 1988) as well as of S-phenylcysteine (McDonald et al., 1994) in vivo after benzene exposure provides indirect evidence for the formation of benzene oxide as the initial benzene oxidation product. Using HPLC analysis, benzene oxide was tentatively identified as such by Lovern et al. (1997) in microsomal systems metabolizing benzene. Direct evidence for the formation of benzene oxide in vivo after benzene administration comes from studies by Lindstrom et al. (1997). These investigators initially spiked blood from F344 rats with benzene oxide and, using a GC–MS method for measuring the remaining benzene oxide, found that benzene oxide has an estimated half-life of 7.9 min. This half-life is considerably longer than that of less than 2 min reported by Jerina and Daly (1974) for benzene oxide in 1 M KCl at 30 C. An even longer half-life of 34 min was recently reported by Henderson et al. (2005) for benzene oxide dissolved in phosphate buffer in D2O containing deuterated dimethylsulfoxide (95:5, v:v) at 25 C at pD 7.0 (pD ¼ pH þ 0.4). This half-life did not significantly change in the presence of 2–15 mM GSH, but did decrease at higher GSH concentrations. The major product was phenol. Addition of glutathione transferase (GST) to a reaction mixture containing 2 mM GSH also did not appreciably affect the half-life of benzene oxide. The authors concluded that capture of benzene oxide (which is in equilibrium with benzene oxepin) by GSH is an inefficient process that may account for the low levels of S-phenylmercapturic acid present in the urine after benzene exposure. Little is known, however, about the stability of benzene oxide/oxepin in hydrophobic environments, such as lipophilic portions of cell membranes and hydrophobic pockets of proteins, in which it may be sequestered and protected from reacting with glutathione. The favorable entropy in such microenvironments may actually promote reaction of benzene oxide with thiols, such as cysteine residues of albumin and hemoglobin, reactions that form the basis of a specific biomarker assay for benzene exposure. Using the GC–MS method developed for benzene oxide in rat blood, Lindstrom et al. (1997) demonstrated that benzene oxide is indeed formed in vivo in F344 rats administered 400 mg/kg benzene. In subsequent studies, Lindstrom et al. (1998) reported similar half-lives for benzene oxide incubated with blood from humans (7.2 min) and mice (6.6 min). Although not formed in microsomal preparations of rat bone marrow, benzene oxide was also found to have an estimated half-life of 6 min in bone marrow homogenates of F344 rats (Lindstrom et al., 1999). Thus, in contrast to what was previously believed, benzene oxide is a relatively stable electrophile, with a half-life long enough to be distributed via the blood stream to the bone marrow and other target tissues. The studies by Lindstrom et al. described above not only definitely demonstrated that benzene oxide is formed in vivo, but also formed the basis for estimating second-order rate constants for the reaction of benzene oxide with cysteinyl residues of hemoglobin and

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albumin, which these investigators subsequently used to assess systemic doses of benzene oxide arising from benzene exposure in previously published animal and human studies. Phenol can undergo further cytochrome P450-mediated oxidation to catechol and hydroquinone. Benzene oxide can also be metabolized by epoxide hydrolase to benzenetrans-dihydrodiol, a metabolite converted to catechol by the action of a dehydrogenase. The 1,2,4-benzenetriol, observed in some in vitro metabolism systems, is believed to be derived from the oxidation of dihydroxylated metabolites by cytochrome P450. The hydroxylated aromatic benzene metabolites can undergo further metabolic conversion to form sulfate or glucuronic acid conjugates. Benzene epoxide can also serve as substrate for glutathioneS-transferase, catalyzing the formation of the prephenylmercapturic acid, which is aromatized by dehydration under acidic conditions to S-phenylmercapturic acid (Sabourin et al., 1988b). Tunek et al. (1980) identified the glutathione conjugate of p-benzoquinone in an incubation of phenol and glutathione with microsomes. This product is thought to be formed mainly nonenzymatically (Lunte and Kissinger, 1983). The cytochrome P450 isozyme primarily responsible for the initial oxidation of benzene is cytochrome P4502E1 (Johansson and Ingelman-Sundberg, 1988; Koop et al., 1989; Schrenk et al., 1992). This P450 isozyme is induced by many chemicals, including ethanol, a chemical known to enhance benzene hematotoxicity in mice when administered in the drinking water (Baarson et al., 1982). In a study on benzene metabolism in relation to cytochrome P4502E1 activity, Seaton et al. (1994) reported that measured cytochrome P4502E1 activities varied 13-fold for microsomes prepared from human liver samples, and that the fraction of benzene metabolized in 16 min ranged from 10% to 59%. A model developed by the investigators predicted the dependence of benzene metabolism on the measured cytochrome P4502E activity in liver samples from humans, rats, and mice. The authors suggested that interindividual and interspecies variations in hepatic metabolism of benzene may be related to differences in liver cytochrome P4502E1 activity. Since benzene metabolism is required for toxicity, the findings suggest that interindividual differences in P4502E1 gene expression, including differences in P4502E1 induction, as a result of alcohol consumption, for example, could play a role in determining susceptibility to benzene toxicity. The relationship between cytochrome P450 expression and benzene metabolism and toxicity is further discussed below. A reactive ring-opened metabolite of benzene trans,trans-muconaldehyde, a reactive ring-opened metabolite of benzene, was identified by Latriano et al. (1986) as a microsomal metabolite of benzene, but it has not been identified in vivo. This metabolite, a six-carbon diene dialdehyde, is unique among individual benzene metabolites in its ability to cause bone marrow depression in mice (Witz et al., 1985). Muconaldehyde was originally identified as a product formed via hydroxyl radical-mediated ring opening in aqueous solutions of benzene irradiated with X-rays (Loeff and Stein, 1959). Studies by Latriano et al. (1985) showed that muconaldehyde is formed from benzene in the presence of a hydroxyl radical-generating Fenton system. In subsequent studies, both the trans,trans- and the cis,trans-isomers of muconaldehyde were identified in Fenton mixtures incubated with benzene (Zhang et al., 1995a). Identification of cis,trans-muconaldehyde, an isomer most likely derived by rearrangement of the less stable cis,cis-muconaldehyde, suggests that cis,cis-muconaldehyde is the initial product of benzene ring opening in a Fenton system. Benzene dihydrodiol, a newly identified product derived from benzene incubated in the Fenton system, was shown to form phenol, catechol and ring-opened a,b-unsaturated aldehydes of unknown structure. These results indicate that although benzene dihydrodiol can be ring opened to a,b-unsaturated aldehydic products, it is not the precursor of muconaldehyde formed from


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benzene in a Fenton system. Studies by Zhang et al. (1995b) suggest that Fenton chemistry plays a role in the formation of ring-opened as well as ring-hydroxylated compounds derived from benzene upon incubation with liver microsomes. Using microsomes prepared from the liver of mice treated with acetone for induction of cytochrome P4502E1, Zhang et al. showed that ring opening and ring hydroxylation were enhanced by micromolar concentrations of iron and inhibited by addition of oxy radical scavengers. Cytochrome P4502E1, the major isozyme responsible for the initial oxidation of benzene to benzene oxide, is known to produce large amounts of hydrogen peroxide (Wu and Cederbaum, 1994; Kukielka and Cederbaum, 1995). The possibility exists that benzene ring opening is mediated by reactive oxygen species generated during P4502E1 metabolism and/or involves a series of steps that include enzymatic and nonenzymatic transformations. Oxidative and reductive metabolism of muconaldehyde leads to a variety of metabolites, some of which could be important in muconaldehyde toxicity, and consequently in benzene hematotoxicity (Witz et al., 1996). Of particular interest is 6-hydroxy-trans,trans2,4-hexadienal, a reduced muconaldehyde metabolite shown to be hematotoxic in mice (Zhang et al., 1995c) and mutagenic in V79 cells (Chang et al., 1994). 6-Hydroxy-trans, trans-2,4-hexadienal is less reactive than muconaldehyde and, if formed in the liver (Grotz et al., 1994), is more likely than muconaldehyde to survive transport to the bone marrow, the target tissue of benzene. The monoreduction of muconaldehyde is reversible (Zhang et al., 1995a), a finding that could relate to the hematotoxicity observed for muconaldehyde as well as for 6-hydroxy-trans,trans-2,4-hexadienal. The role of free radicals in benzene metabolism and benzene toxicity is relatively unexplored (Subrahmanyam et al., 1991). Using a reconstituted system containing rabbit liver P450 isozyme LM2 as well as microsomes, Johansson and Ingelman-Sundberg (1983) demonstrated that phenol formation from benzene is inhibited by presumed hydroxyl radical scavengers including mannitol and dimethyl sulfoxide (DMSO) and by catalase, horseradish peroxidase, and superoxide dismutase. The authors suggested that the cytochrome P450dependent metabolism of benzene to phenol is mediated by hydroxyl radicals generated from hydrogen peroxide, thought to be formed by the spontaneous dismutation of superoxide anion radicals released by cytochrome P450. In this mechanism of phenol formation, the initial reactive intermediate is a hydroxy cyclohexadienyl radical formed by the addition of a hydroxyl radical to the benzene ring. This reaction takes place readily (Walling and Johnson, 1975), followed by phenol formation via loss of a hydrogen atom. In a subsequent study on the role of free hydroxyl radicals in the cytochrome P450-catalyzed oxidation of benzene and cyclohexanol, Gorsky and Coon (1985) concluded that, in the presence of very low (micromolar) concentrations of benzene, the hydroxyl radical-mediated formation of phenol is the dominant pathway, whereas at higher concentrations (millimolar), the direct oxidation by P450 is quantitatively of much greater importance. One implication of these studies is that a shift in metabolic pathways (free radical-mediated compared with direct enzymatic conversion) may occur, depending on the exposure dose or concentration of benzene. More recent approaches to evaluating the potential role of oxidative stress in benzene toxicity have utilized toxicogenomics (Hirabayashi, 2005). The bone marrow, the target tissue of benzene, has been reported to contain small amounts of cytochrome P450 (Andrews et al., 1979) and to metabolize benzene to only a limited extent (Irons et al., 1980). Cytochrome P4502E1 was not detected in bone marrow of B6C3F1 mice (Genter and Recio, 1994), but it is not known whether other strains of mice or other species also lack this major benzene metabolizing P450 isozyme in the bone marrow. However, the hydroxylated metabolites of benzene are present in the bone marrow after

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benzene administration, and the levels of catechol and hydroquinone persist in this tissue (Rickert et al., 1979; Irons et al., 1982), suggesting a role for these metabolites in benzene toxicity. Peroxidases are present in the bone marrow, and recent studies suggest that they may play a role in the metabolism of hydroxylated benzene metabolites to reactive toxic intermediates. Using horseradish peroxidase (HRP) in the presence of hydrogen peroxide as a model for bone marrow peroxidases, Sawahata and Neal (1982) demonstrated the formation of biphenols and of p-diphenoquinone from phenol. The formation of biphenols and covalent binding of 14 C to protein was also observed during the incubation of [14 C] phenol with a rat bone marrow homogenate in the presence of hydrogen peroxide. The peroxidative oxidation of phenol and other hydroxylated benzene metabolites to reactive intermediates and the ability of hydroxylated benzene metabolites to serve as good reducing cosubstrates for peroxidases in the oxidation of benzene metabolites to reactive intermediates are well documented (Subrahmanyam and O’Brien, 1985; Smart and Zannoni, 1985; Eastmond et al., 1987; Sadler et al., 1988). The reactive toxic intermediates generated by metabolism of hydroxylated benzene metabolites by myeloperoxidase, the major peroxidase in the bone marrow, are quinones and free radical metabolite intermediates. The quinones can be detoxified by a two-electron reduction by NAD(P)H:quinone acceptor oxidoreductase (NQO1), a process that regenerates the polyhydroxylated benzene metabolites. The balance of activation of hydroxylated benzene metabolites and detoxification by NQO1 has been suggested to be a determining factor in benzene bone marrow toxicity (Ross, 2000). Results from a study by Rothman et al. (1997) support the hypothesis that high cytochrome P4502E1 activity along with low NQO1 activity are susceptibility factors for benzeneinduced hematotoxicity. The major route of human exposure to benzene is via inhalation; consequently, the lung is the first site for absorption and potential metabolism of benzene. Little is known about lung metabolism of benzene in humans compared with experimental animals. A pharmacokinetic modeling study by Sherwood and Sinclair (1999) suggested that organs other than the liver may contribute significantly to benzene metabolism, and the lung certainly is an important organ to consider. Initial studies by Chaney and Carlson (1995) showed that rat pulmonary microsomes metabolize benzene, and later studies by Powley and Carlson (1999), on species comparisons of humans, mice, rabbits, and rats, demonstrated that the rat is most similar quantitatively and qualitatively to human pulmonary microsomal metabolism of benzene incubated at 24–1000 mM concentrations. In the same studies, the rat was also found to be similar to the human in the metabolism of low concentrations of benzene (24 and 200 mM) by hepatic microsomes, while benzene metabolism to phenol by mouse microsomes is most similar to human at higher benzene concentrations (700 and 1000 M). Interestingly, hepatic, but not pulmonary, microsomes exhibited saturation of benzene metabolism, and a greater proportion of phenol was converted to hydroquinone when the benzene concentration was increased in pulmonary microsomes, while the opposite was observed with hepatic microsomes. The authors concluded that overall the rat is most similar to the human in oxidative benzene metabolism at the lower environmentally relevant benzene levels. In subsequent studies using microsomal preparations from CYP2E1 knockout and wild-type mice, Powley and Carlson (2000) showed that, in contrast to the liver where CYP2E1 was the most important isozyme accounting for 96% of total hydroxylated metabolite formation, lung CYP2E1 was responsible for only 45% of total hydroxylated metabolite formation and that CYP2F2 also contributed to the formation of oxidized benzene metabolites. Additional studies suggest that CYP2E1 is less important in the lung than the liver and show that it has a lower affinity for benzene, but a higher rate of hydroxylated metabolite formation than


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CYP2F2, which plays a predominant role in benzene metabolism in the mouse lung. Recent studies of benzene metabolism using human bronchiolar- and alveolar-derived cell lines (Sheets and Carlson, 2004) demonstrate that benzene is metabolized by CYP2E1 and CYP2F1 that are expressed in the lung. 13.3.3

Species and Strain Differences in the Metabolism of Benzene

Chronic toxicity studies by the National Toxicology Program (NTP) (Huff, 1983) have shown that B6C3F1 mice are more sensitive to the hematotoxic and carcinogenic effects of benzene than F344/N rats. The greater sensitivity of mice compared with rats to benzene toxicity may be related to a greater metabolism of benzene to toxic intermediates, differences in detoxification of toxic metabolites, or greater inherent susceptibility of target tissues to the action of toxic metabolites. Differential toxicity as related to metabolic differences between B6C3F1 mice and F344/N rats was investigated when Sabourin et al. (1988a) quantitated water-soluble metabolites for four metabolic pathways. In animals exposed to 50 ppm benzene for 6 h, phenylsulfate, a detoxification metabolite, was present in approximately equal concentrations in rats and mice. Hydroquinone glucuronide, hydroquinone, and muconic acid, which are thought to reflect pathways leading to potential toxic metabolites of benzene, were present in much greater concentration in the mouse than in the rat. These results suggest that greater metabolism to toxic intermediates may in part explain the higher susceptibility of B6C3F1 mice to benzene-induced toxicity. Significant differences in the metabolism of benzene to urinary muconic acid have also been demonstrated between DBA/2N and C57BL/6 mice (Witz et al., 1990), two strains previously reported to exhibit differences in benzene toxicity (Longacre et al., 1981). At hematotoxic benzene doses (220–880 mg/kg), benzene-sensitive DBA/2N mice excreted significantly more muconic acid than the less-benzene-sensitive C57BL/6 mice. No differences between the two strains were observed in urinary excretion of muconic acid after muconaldehyde administration or muconic acid administration. Assuming that urinary muconic acid is derived from muconaldehyde (Kirley et al., 1989), these findings suggest that strain sensitivity toward benzene may be related to differences in the metabolism of benzene to toxic intermediates, including toxic ring-opened compounds such as muconaldehyde. In contrast to the results at hematotoxic benzene doses, at lower benzene doses (0.5–2.5 mg/kg) C57BL/6 mice excreted significantly more muconic acid than DBA/2N mice. These findings may reflect a dose-dependent (and strain-specific) shift in the metabolic pathways of benzene analogous to that found by Gorsky and Coon (1985) for the in vitro metabolism of benzene. The role of benzene metabolism, in relation to toxicity, was recently investigated by Valentine et al. (1996) in male mice lacking cytochrome P4502E1 expression (Lee et al., 1996). The CYP2E1 knockout mice and their wild-type counterparts were F3 homozygous hybrids of SV/129 C57BL/6N. After a 6 h nose-only exposure to 200 ppm benzene along with a tracer dose of [14 C]benzene, total urinary metabolites in the knockout mice were decreased to 13% compared with total urinary metabolites excreted by wild-type control mice. The amount of phenylsulfate excreted in the knockout mice constituted a significantly larger percentage of urinary total radioactivity than that in the wild-type mice, indicating a substantial role for cytochrome P4502E1 in the oxidation of phenol formed from benzene in normal mice. Toxicity studies in the knockout mice exposed by whole body inhalation to 0 or 200 ppm benzene 6 h/day for 5 days showed no effects on bone marrow cellularity and no benzene-induced genotoxicity using micronuclei formation in bone marrow and blood as an

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end point. In contrast, wild-type and B6C3F1 mice exhibited severe bone marrow cytotoxicity and genotoxic effects in bone marrow and blood. These studies indicate that cytochrome P4502E1 is the major benzene-metabolizing P450 isozyme in vivo and that benzene metabolism is required for toxicity. These conclusions are supported by other in vivo studies with male and female B6C3F1 mice, which exhibited sex-dependent differences in the rate of benzene metabolism that are correlated with known differences in genotoxicity (Kreyon et al., 1996). 13.3.4

Disposition of Benzene

In humans, the half-life of benzene is in the order of 1–2 days, and essentially all absorbed benzene is gone from the body within a week following exposure. Thus, long-term bioaccumulation of benzene or its metabolites is not of concern. Ideally, in order to extrapolate results from animal experiments to humans, investigators should know the effects of dose, exposure rate, route of exposure, and species on the metabolism and disposition of benzene. Sabourin et al. (1987) found virtually 100% absorption in F344/N and Sprague–Dawley rats and B6C3F1 mice that had received oral administration of 0.5–150 mg/kg benzene. This differs from inhalation, where the percentage of benzene absorbed and retained during a 6 h exposure decreases as the exposure concentration increases. For example, at 10 ppm benzene, mice and rats absorb and retain 33% and 50%, respectively, compared with 15% and 10% respectively, at 1000 ppm benzene. At oral doses below 15 mg/kg benzene, mice and rats excreted more than 90% of the administered dose as urinary metabolites. Above 15 mg/kg, increasing amounts of benzene were exhaled unmetabolized with increasing exposure concentration, suggesting saturation of metabolism of orally administered benzene in rats and mice. For inhalation exposures, saturation of metabolic routes occurred in mice, but not in rats, at higher exposure concentrations. The results indicate that a saturating dose, if given as a bolus by gavage, is not saturating when administered by inhalation over 6 h. The effect of exposure concentration, exposure rate, and route of administration on metabolism was studied by Sabourin et al. (1989) in F344/N rats and B6C3F1 mice. Animals were exposed orally to 1, 10, and 200 mg/kg benzene and by inhalation for 6 h to 5, 50, and 600 ppm benzene vapor. In addition, animals were exposed over different time intervals to the same total amount of benzene (C T ¼ 300 ppm h). As the exposure concentration or oral dose increased, there was a shift in metabolism from putative toxification pathways to detoxification pathways. In mice, hydroquinone glucuronide and muconic acid (markers of toxification pathways) represented a greater percentage of the administered dose at low benzene doses than was evidenced at high doses. The percentage dose excreted as the detoxification products phenylglucuronide and prephenylmercapturic acid increased with increasing dose. Similar results were obtained in the rat, except that hydroquinone glucuronide was a minor benzene metabolite at all concentrations. No simple relationship between oral dosing and inhalation was observed in terms of metabolite dose to tissues. These studies indicate that extrapolation from high-exposure toxicity studies to low-level exposures or from oral to inhalation exposures may not result in a true estimate for the parameter being extrapolated to humans. The authors concluded that if hydroquinone glucuronide and muconic acid are markers of toxic benzene metabolites and are formed in significant amounts in humans, then linear extrapolation of health effects from high exposures to low-level exposures could underestimate the toxicity of benzene. These data as well as the previous studies of Witz et al. (1990) and recent work by Rappaport et al. (2005)

MECHANISMS OF TOXICITY

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all suggest the possibility of a hyperbolic dose–response curve for benzene hematotoxicity and carcinogenesis

13.4 MECHANISMS OF TOXICITY Remaining as major uncertainties are the nature of the toxic metabolites responsible for benzene toxicity and the mechanisms of action of these toxic metabolites. Any hypothesis of benzene toxicity must account for the role of hepatic metabolism and the selective toxicity of benzene in the bone marrow. A longstanding puzzling aspect of benzene toxicology is its lack of hepatotoxicity. Adding to this puzzle is the finding by Heijne et al. (2005) that hepatic gene expression in Fisher rats exposed to benzene is similar to that observed following exposure to known hepatotoxins. As discussed above, benzene is metabolized in the liver, mainly to phenol and minor amounts of hydroquinone and catechol. In contrast to muconaldehyde, the hydroxylated benzene metabolites administered singly do not cause bone marrow toxicity in experimental animals. Phenol and hydroquinone, which accumulate in bone marrow (Greenlee et al., 1981), have been shown to cause significant decreases in bone marrow cellularity when coadministered to mice (Eastmond et al., 1987). Since the bone marrow is rich in peroxidative enzymes, including myeloperoxidase and potential oxidants such as hydrogen peroxide derived from leukocytes, it has been suggested (Eastmond et al., 1987; Smith et al., 1989) that a bone marrow-localized phenol-dependent stimulation of hydroquinone metabolism results in the formation of benzoquinone, the ultimate toxic benzene metabolite. Benzoquinone is a direct-acting alkylating agent. It readily reacts with sulfhydryls and has been shown to inhibit microtubule assembly by blocking the thiol-sensitive GTP binding site (Irons et al., 1981). Benzoquinone also forms DNA adducts (Jowa et al., 1990), causes DNA strand breakage (Pellack-Walker and Blumer, 1986), and is genotoxic in V79 cells (Glatt et al., 1989). Thus, binding of benzoquinone to critical cellular substituents may play a role in benzene myelotoxicity. Recent work by Gaskell et al. (2005) has explored the formation of DNA adducts from benzene and the benzene metabolites para-benzoquinone and hydroquinone. The bone marrow is a complex matrix harboring stem cells, progenitor cells of blood cells, and stromal cells, which provide growth factors necessary for the proliferation and differentiation of stem and progenitor cells (Tavassoli and Friedenstein, 1983). The stromal macrophage, a regulator of hematopoiesis (Bagby, 1987), has been proposed to be a specific target of benzene (Kalf et al., 1989). In DBA/2N and C57BL/6 mice, benzene caused a dose-dependent bone marrow depression and a significant increase in bone marrow prostaglandin E levels. Both effects were prevented by coadministration of indomethacin and other inhibitors of the cyclooxygenase component of prostaglandin H synthase (PHS). Benzene, or a reactive metabolite, is hypothesized to stimulate the release of arachidonic acid, which is further metabolized to the hydroperoxide PGG2. In this mechanism of benzene toxicity, the decreased bone marrow cellularity after benzene administration is attributed to the constitutive production of high levels of prostaglandins, known to be downregulators of hematopoiesis, coupled with the genotoxic damage from reactive metabolites such as benzoquinone. Using hydroquinone or phenol as electron donors, the endoperoxidase would metabolize PGG2 to PGH2, the immediate precursor molecule for prostaglandins, with the concomitant formation of benzoquinone. Phagocytic cells have the ability to produce a variety of toxic oxygen species including superoxide anion radical, hydrogen peroxide, and hydroxyl radical. Bone marrow

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macrophages and granulocytes from Balb/c mice treated with 880 mg/kg benzene were found to produce elevated levels of hydrogen peroxide on stimulation with phorbol myristate acetate compared with the same types of cells from control mice (Laskin et al., 1989). In addition to toxic oxygen species, Laskin et al. (1995) demonstrated that bone marrow leukocytes from mice administered hematotoxic doses of benzene or the metabolites hydroquinone, p-benzoquinone, and 1,2,4-benzenetriol produced increased amounts of nitric oxide (NO) in response to the inflammatory mediators lipopolysaccharide or interferon gamma. The production of NO induced by the inflammatory mediators was further enhanced by granulocyte macrophage and macrophage colony-stimulating factor, that is, growth factors present in the bone marrow required for normal cell proliferation and differentiation. The authors suggest that elevated NO production in the bone marrow may be an important mediator of benzene-induced bone marrow suppression (Laskin et al., 2000). Further supporting a role for elevated bone marrow production of NO as an important mediator of bone marrow effects is the work of Chen et al. (2004) who identified nitrobenzene, nitrobiphenyl, and nitrophenol in bone marrow of mice after i.p. administration of 400 mg/kg benzene. Nitrated benzene metabolites were either not detected in liver, lung, and blood or were present at levels significantly less than that found in the bone marrow. The authors hypothesized that peroxynitrite and other NO-derived intermediates are formed in the bone marrow via reaction of NO with oxygen or superoxide anion generated by redox cycling of hydroxylated benzene metabolites. Using 3-nitrotyrosine as a biomarker for NO-induced damage to proteins, Chen et al. (2005) demonstrated that 3-nitrotyrosine contents in bone marrow proteins increased by 1.5–4.5-fold in B6C3F1 mice administered 50–200 mg/kg benzene compared with controls. At 400 mg/kg benzene, 3-nitrotyrosine content of bone marrow proteins was significantly lower than that observed at 200 mg/kg benzene, but still significantly higher than that of controls. The authors suggest that nitration of proteins by peroxynitrite and/or by bone marrow myeloperoxidase-dependent pathways in NO metabolism may account in part for the myelotoxicity and leukemogenic effects of benzene. Nitration of tyrosine residues of specific proteins could affect signal transduction pathways and could inactivate enzymes in the bone marrow. Whether nitrated benzene metabolites play a role in benzene toxicity is at present not known. Despite many interesting hypotheses and fruitful lines of investigation, it is not now possible to identify a specific metabolic pathway leading to a specific toxic intermediate producing a specific pathogenetic mechanism resulting in either bone marrow aplasia or leukemogenesis. In fact, it has become more apparent that the effect of benzene is likely to be exerted through the action of multiple metabolites on multiple end points through multiple biological pathways (Goldstein, 1989b; Eastmond et al., 1987). An overall hypothesis for benzene-induced leukemia was proposed by Smith (1996). The key elements of this hypothesis consist of generally accepted knowledge on benzene metabolism and disposition of metabolites, cellular targets, and effects leading to changes in structure, which result in protooncogene activation and the inactivation of tumor suppressor genes. A stem cell thus affected would proliferate and develop a leukemic clone, which then develops into the disease state. Among the molecular targets of toxic metabolites suggested by Smith are tubulin, DNA, and topoisomerase II (topo II). The inhibitory effect of p-benzoquinone on tubulin formation has been known for some time (Irons et al., 1981). In addition to DNA adduct formation, which could lead to mutations and cancer discussed above, Kolachana et al. (1993) showed that mice administered benzene or its phenolic metabolites have increased levels of 8-hydroxydeoxyguanosine (8-OHdG) in bone marrow cell DNA. A study by Lagorio et al. (1994) in 65 filling station attendants in Rome, Italy,


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showed a dose–response relationship between personal exposure to benzene and urinary concentrations of 8-OHdG. Studies by Frantz et al. (1996) and Chen and Eastmond (1995a) indicate that topoisomerase II is inhibited directly by p-benzoquinone and muconaldehyde, and by phenol and the polyhydroxylated metabolites after their peroxidase activation. Work by these groups and others has continued to explore the role of topoisomerase in benzene hematotoxicity (Lindsey et al., 2005). Topo II is involved in breaking and resealing DNA strands during DNA replication and repair and inhibition of topo II could lead to chromosome breaks and aneuploidy or cell death. It is of interest to note that the use of topo II inhibitors in chemotherapy is associated with a high risk of developing acute myeloid leukemia (Francis et al., 1994). The inhibition of topo II by benzene metabolites could contribute to the clastogenic and carcinogenic effects of benzene. Synergistic clastogenic effects have been observed in bone marrow erythrocytes of mice coadministered phenol and hydroquinone (Barale et al., 1990). The increase in micronuclei induction in the bone marrow erythrocytes appears to originate mainly from breakage in the euchromatic region of mouse chromosomes (Chen and Eastmond, 1995b). Whysner et al. (2004) and Eastmond et al. (2005) have recently reviewed the evidence concerning benzene genotoxicity and have concluded that a role for topoisomerase is likely. Elevated levels of DNA–protein cross-links (DNAPC) have been demonstrated in bone marrow cells of mice administered benzene and in HL-60 cells exposed to muconaldehyde (Schoenfeld et al., 1996). Structure–activity relationship studies of muconaldehyde and its metabolites in HL-60 cells (Schoenfeld and Witz, 2000) indicate that 6-hydroxy-trans, trans-2,4-hexadienal, the initial reduction product, and 6-oxo-trans,trans-2,4-hexadienoic acid, the initial oxidation product of muconaldehyde, also have the ability to induce DNA–protein cross-links, albeit at 5–10 times higher concentrations than muconaldehyde. DNA–protein cross-links could be involved in the hematotoxicity of 6-hydroxy-trans, trans-2,4-hexadienal, as well as that of muconaldehyde and benzene. Hydroquinone was subsequently also shown to induce DNAPC, but to a considerably lesser extent than muconaldehyde (Amin and Witz, 2001). DNAPC formation by hydroquinone could potentially be mediated by 1,4-benzoquinone, its oxidation product, or by reactive oxygen species generated during hydroquinone metabolism. Incubation of HL-60 cells with equimolar mixtures of muconaldehyde and hydroquinone resulted in higher DNAPC levels relative to those expected if the effects were additive. The induction of DNA–protein cross-links by toxicants is often associated with DNA strand breaks (Cosma et al., 1988; Yamanaka et al., 1995), suggesting that DNA–protein cross-links could contribute to the observed clastogenic effects observed after benzene exposure. Studies by Amin and Witz (2001) indicate that both muconaldehyde and hydroquinone cause concentration- and time-dependent increases in DNA single- and double-strand breaks (DNASB) and alkalilabile sites in HL-60 cells as determined by a fluorometric assay based on DNA unwinding. Induction of DNASB was additive upon treatment with equimolar mixtures of muconaldehyde and hydroquinone. 30 OH DNASB levels determined by the TUNEL assay increased significantly in HL-60 cells in a concentration-dependent manner after treatment with muconaldehyde (5–25 mM, 1 h), while hydroquinone had no effect. Cotreatment with equimolar muconaldehyde/hydroquinone mixtures resulted in significant decreases in 30 OH DNASB compared to treatment with muconaldehyde without hydroquinone. If the 30 OH DNAPC caused by muconaldehyde are an indication that apoptosis has been induced, the results suggest that a muconaldehyde/hydroquinone interaction could potentially result in survival of cells with genotoxic damage. It would be of interest to pursue this interaction in bone marrow cells and study the effects of

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mixtures of these and other benzene metabolites on signal transduction pathways involved in apoptosis. In vitro studies in bone marrow cells and HL-60 cells indicate induction of apoptosis by exposure to phenolic metabolites and to muconaldehyde, suggesting that programmed cell death could be involved in benzene bone marrow cytotoxicity (Moran et al., 1996; Hiraku and Kawanishi, 1996; Schoenfeld et al., 1997). Studies analyzing bone marrow cell populations in mice administered benzene or hydroquinone (Hazel et al., 1996) and in vitro studies in model cell systems with reactive benzene metabolites (Irons and Stillman, 1996; Hedli et al., 1996; Hazel and Kalf, 1996; Kalf et al., 1996) indicate changes in cell differentiation and/or effects on signal transduction pathways, which could play a role in the mechanism of benzene-induced leukemogenesis. Similarly, a possible role for benzene metabolites, and particularly muconaldehyde, in leukemogenesis through interference with gap junction intercellular communication has been suggested (Rivedal and Witz, 2005). One particularly exciting area of benzene research can be loosely grouped under the heading of biological markers (NAS, 1989). As more is learned about the metabolism of benzene and the pathogenesis of its effects, it becomes possible to develop markers of exposure or effect suitable for assay in body fluids of laboratory animals or of humans. These should be of value in elucidating the mechanism of benzene hematotoxicity, in determining the extent of human exposure to benzene, and in establishing the appropriate dose–response curve for the effects of benzene. Substantial efforts have been made in the past 10 years toward the development of biological markers or biomarkers of benzene exposure and their application to exposed populations. Biomarkers have the advantage over external monitoring of integrating exposure from all routes and sources. Their measurement quantifies internal dose and reflects metabolism, where applicable, and disposition. Potential biomarkers of benzene exposure are the parent compound benzene, ring-hydroxylated and ringopened metabolites, glutathione-derived metabolite adducts, and DNA- and protein-derived adducts. Analytical methods for measurement of urinary biomarkers based on benzene metabolites and protein adducts based on cysteinal adducts of benzoquinone and benzene oxide were developed in the 1990s and the sensitivity and specificity of these biomarker assays have been described in several reviews (Bechtold and Henderson, 1993; Ong et al., 1995; Medeiros et al., 1997). Improvement of these methods in recent years has led to the development of less cumbersome and more sensitive assays and their application has resulted in significant progress not only in measuring benzene exposure, but also in other important aspects of benzene toxicology including susceptibility factors for benzene toxicity and correlations with toxic effects. For a detailed description, the reader is referred to articles published in a Special Issue of Chemico-Biological Interactions (Vol. 153–154, May 2005) from the Proceedings of the International Symposium on Recent Advances in Benzene Toxicity, held in Munich, Germany, October 9–12, 2004. Among the highlights of recent biomarker studies is the finding of a significant correlation of 1,4-benzoquinone (1,4-BQ) adducts of albumin (Alb) and hemoglobin (Hb) in Chinese workers exposed to benzene (Yeowell-O’Connell et al., 2001). In this study, 1,4-Alb adducts and albumin adducts of benzene oxide (BO) were found to be highly correlated with each other and with urinary phenol (PH) and hydroquinone (HQ) when compared on an individual basis. Another study in Chinese workers (Qu et al., 2003) indicated that the urinary biomarkers S-phenylmercapturic acid (S-PMA) and trans,trans-muconic acid (t,t-MA) showed significant exposure–response trends even at low benzene levels, and that S-PMA, which detects benzene exposure of about 0.1 ppm, is superior to t,t-MA, which detects benzene levels of 1.0 ppm. The hydroxylated metabolites HQ, catechol (CAT), and PH were


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only able to detect exposures above 5 ppm. Measurement of 1,4-BQ and BO albumin adducts in 134 workers exposed to benzene (0.07–46.6 ppm; median 3.55 ppm) and 51 unexposed controls in Tianjin, China, provided strong evidence that metabolism of benzene by CYP2E1 is saturated at occupational exposures of about 1 ppm (Rappaport et al., 2002). This is based on adduct levels that were not linear and less than proportional with benzene exposure. The 1,4-BQ/BO adduct ratio was found to decrease with age, coexposure to toluene, and alcohol consumption, findings interpreted to indicate that factors affecting CYP2E1 exert a greater effect on the production of 1,4-BQ than that of BO, most likely due to the second oxidation step required for the formation of HQ. In a subsequent study (Waidyanatha et al., 2004) on benzene-exposed workers (1.65–329 ppm) and 44 controls in Shanghai, China, a similar conclusion of saturation of benzene metabolism was reached based on urinary biomarkers. For these studies, a newly developed method based on solvent extraction of 0.5 mL acidified urine, trimethylsilyl derivatization, and GC–MS analysis was used to measure PH, CAT, HQ, 1,2,4-trihydroxybenzene (triOHBz), t,t-MA, and S-PMA. There was a greater than proportional production of PH, CAT, and S-PMA and a less than proportional production of HQ, t,t-MA, and triOHBz, findings that are consistent with a competitive inhibition of PH, BO, and HQ for the same CYP2E1 enzymes. HQ, t,t-MA, and triOHBz all require more than one oxidation step for their formation, and their production is expected to be decreased under conditions of competitive inhibition of CYP2E1. The fact that CATwas grouped with PH and S-PMA was interpreted to suggest that catechol is formed in humans from the dihydrodiol obtained by ring opening of benzene oxide, and not via a second oxidation of phenol. Additional studies on cysteinyl albumin adducts of 1,4-BQ and BO as biomarkers of human benzene metabolism in Chinese workers by Rappaport et al. (2005) support the above results, indicating that adduct formation is less than proportional to benzene exposure at above about 1 ppm. The authors suggest that the biologically effective dose of 1,4-BQ and BO should be proportionally greater in persons exposed to low rather than high levels of benzene. This could have profound implications with respect to risk assessment, as discussed later in this chapter. Since the toxicity of benzene requires metabolism, factors that affect enzymes involved in the activation of benzene and its metabolites to toxic intermediates as well as enzymes involved in the deactivation of toxic metabolites may modulate benzene toxicity by changing the levels of metabolites produced. In recent years, increasing use has been made of genetically altered animal models to probe not only the role of key enzymes in the hematotoxic and genotoxic effects of benzene (Long II et al., 2002; Bauer et al., 2003) but also regulatory mechanisms in signal transduction that could be dysregulated and involved in myelotoxicity (Boley et al., 2002; Nwosu et al., 2004). In human populations, studies have focused on polymorphisms in metabolic genes in relation to biomarkers of benzene exposure and correlation with increased susceptibility to benzene toxicity. Key enzymes studied in human populations include CYP2E1 and NAD(P)H quinone:oxidoreductase (NQO1), that is, the major enzyme involved in the initial activation of benzene to benzene oxide in the liver and a major enzyme involved in the deactivation of toxic quinone metabolites in the bone marrow, respectively, as well as myeoloperoxidase (MPO), which is involved in the metabolism of phenolic metabolites to toxic quinones in the bone marrow and glutathione transferase GSTT1, which catalyzes glutathione adduct formation with benzene oxide, thereby deactivating benzene oxide, the initial reactive oxidation product. A detailed account of these studies is beyond the scope of this review, and the reader is referred to the Special Issue of Chemico-Biological Interactions (Vol. 153–154, 2005) mentioned earlier. A highlight among these studies is the finding of increased susceptibility to benzene toxicity

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in Chinese workers with a polymorphism in NQO1 consisting of a C to T point mutation at position 609 in exon 6. This mutation codes for a proline-to-serine change at position 187 in the protein, resulting in instability and rapid degradation. Individuals with two null alleles do not exhibit NQO1 activity. In a study by Rothman et al. (1997), NQO1 and CYP2E1 genotypes were determined and CYP2E1 activity was estimated from the fractional excretion of hydroxychlorzoxazone. A 7.6-fold increased risk of benzene poisoning was observed in rapid chlorzoxazone metabolizers who were homozygous for the NQO1 polymorphism compared with slow chlorzoxazone metabolizers who had two wild-type NQO1 alleles. The PstI/RsaI polymorphism was found not to influence the risk of benzene poisoning. In an NQO1 mouse animal model, benzene exposure was shown to decrease apoptosis and cause myelogenous hyperplasia in the bone marrow and significant increases in peripheral blood neutrophils, eosinophils, and basophils (Long II et al., 2002). These studies support a critical role of NQO1 in myelotoxicity and the authors suggest that NQO1 null mice may potentially be useful as an animal model for studying acute leukemias and chemotherapeutic agents against these diseases. For a detailed account of the NQO1 in relation to benzene myelotoxicity, the reader is referred to a review by Ross (2005). A direct correlation between a gene polymorphism and a biomarker was reported by Qu et al. (2005) in studies on urinary biomarkers and polymorphisms of several metabolic genes in Chinese workers. Subjects with GSTT1 null alleles excreted significantly less S-PMA than those with wild-type GSTT1. The authors concluded that GSTT1 plays a critical role in determining interindividual variation of S-PMA formation from benzene, and that it is important that GSTT1 genotype be known and taken into account when S-PMA is used as a marker to estimate personal exposure levels. A future goal is continued emphasis on studies on the correlation of biomarkers with markers of susceptibility and effect for elucidation of the processes involved in benzene-induced hematotoxicity and leukemogenesis. Many of the studies of human susceptibility described above depend upon modern advances in technology, grouped under the heading of ‘omics’ that are being applied to understanding the mechanism of benzene toxicity. This topic has recently been reviewed by Smith et al. (2005). Narrowing down the genes and proteins primarily involved in benzene toxicity is a promising avenue (see, for example, studies exploring the role of the oncogenes c-MYB (Wan et al., 2005) and p53 (Yoon et al., 2003; Hirabayashi, 2005). One of the more promising experimental designs to come out of advances in the microarray analysis of gene expression is that of studying pairs of individuals: one exposed and one unexposed control. This approach is based upon the strengths of classic epidemiological case–control analyses, in which a careful match of two individuals on all but one characteristic allows exploration of the impact of exposure. Forrest et al. (2005) used this technique in their microarray analysis of peripheral blood mononuclear cell gene expression from paired benzene-exposed and control workers. RNA from these cells obtained in the field from six exposed–control pairs was subjected to study using cDNA microarrays. Realtime polymerase chain reaction technique was used to follow up on selected genes of interest of which four genes were particularly prominent in differentiating between the benzeneexposed and control subjects. In further studies in benzene-exposed workers by this group, Shen et al. (2006) reported that polymorphisms involved in genes repairing DNA doublestrand breaks increased susceptibility to benzene hematotoxicity, and Lan et al. (2005) found that SNPs could be identified from cytokine, chemokine, and cell adhesion pathways involved in hematopoiesis that could influence benzene hematotoxicity. This is an exciting area that will impact both the risk assessment of benzene and the use of toxicogenomics in risk assessment.


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Hematological Effects

13.4.1.1 Pancytopenic Effects Benzene was first identified as a hematological toxicant in the nineteenth century. Experience since that time has amply confirmed the ability of benzene to destroy bone marrow precursor cells that are responsible for the production of mature circulating blood cells in humans. Similar effects are noted in the many species of laboratory animals that have been experimentally exposed to benzene. Human blood cells are generally considered to be of three types: red blood cells, whose primary function is to deliver oxygen to the tissues; white blood cells, which are involved in the body’s defenses against infection; and platelets, which participate in normal blood coagulation. Exposure to benzene affects the formation of each of these cell types. Severe benzene toxicity produces aplastic anemia, which consists of a marked decrease in the cellularity of the bone marrow and highly significant decrements in circulating red blood cells (anemia), white blood cells (leukopenia), and platelets (thrombocytopenia). Aplastic anemia is a frequently fatal disorder with death usually occurring from infection or hemorrhage. The normal bone marrow has ample reserves. For example, under certain circumstances, six times as many red blood cells as normal can be made. Thus, relatively low levels of a toxicant may decrease bone marrow reserve without causing any clinically recognizable decrease in blood counts. As toxicant levels get higher, initially one can see a decrease in blood counts within the normal range, a seemingly selective fall in one of the three blood cell types, then a mild decrease in each, known as pancytopenia, followed by full-blown aplastic anemia. Benzene produces its pancytopenic and aplastic effects through damage to precursors within the marrow by its metabolites. The earliest bone marrow precursor is a pluripotential stem cell that can mature into precursors of red blood cells (erythoblastic cell line), platelets (megakaryocytic cell line), and granulocytic white blood cells (myelocytic cell line, resulting in polymorphonuclear leukocytes, basophils, and eosinophils). The pluripotential cell appears also to be a source of lymphocytic white blood cells. Circulating lymphocytes are relatively susceptible to benzene in laboratory animals (Wiedra et al., 1981; Snyder et al., 1978; Rozen and Snyder, 1985) and in humans (Goldstein, 1988; Rothman et al., 1996). As with other aplastic agents, an increase in red blood cell mean corpuscular volume is also noted. A particularly important recent study in Chinese workers suggests that careful observation of benzene exposure levels and blood counts will show a hematotoxic effect of less than 1 ppm benzene, a long-used benzene workplace standard. Lan et al. (2004) studied 250 workers exposed to benzene and 140 controls and found that platelet counts and white blood cell counts were significantly lower in the benzene-exposed group, even for exposure below 1 ppm. There was a declining trend in other blood counts with higher exposures. The data for red blood cell mean corpuscular volume, an early sign of benzene hematotoxicity, were not presented. Of note is that the authors also evaluated the level of circulating bone marrow progenitor cells in 9 benzene-exposed and 24 control workers. In keeping with the hypothesis that the decline in circulating blood counts reflected an effect of benzene on progenitor cells, they reported a statistically significant lower level of colony-forming units (CFU) for granulocyte macrophages (CFU-GM), for the granulocyte, erythroid, macrophage, megakaryocyte (CFU-GEMM), and erythroid burst-forming units (BFU-E). Similar observations have been made in the blood and bone marrow of laboratory animals exposed to benzene. However, these findings are in contrast with the

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recent report of Quitt et al. (2004), who studied 24 petroleum refinery workers exposed to 0.28–0.41 ppm benzene and 17 age-matched working relatives. They reported an increase, not a decrease, in CFU-GM and BFU-E and noted an interaction with smoking. It is important that these two very different results be reconciled. Further replication of the Lan et al. (2004) finding of a decrease in blood counts at low levels of benzene exposure would also be of value. A variety of potential mechanisms have been suggested to account for presumed differences in individual susceptibility to the pancytopenic effects of benzene, including such factors as gender, increased bone marrow turnover, and various metabolic factors (see discussion below and Goldstein, 1988). A landmark study is the observation by Rothman et al. (1997) that those individuals with an elevation in both the activity of cytochrome P4502E1, as measured by rapid fractional excretion of chlorzoxazone, and gene mutations, reflecting a reduction in activity of NQO1, were associated with a 7.6-fold greater risk of benzene-induced pancytopenia. As discussed above, CYT P4502E1 is a major activator of benzene metabolism and NQO1 is a presumed detoxifying enzyme of benzene metabolites. Either variation by itself showed a two- to threefold increase in risk. Lan et al. (2004), in the study described above, evaluated four different single-nucleotide polymorphisms of potential significance to benzene toxicity. They found that genetic variations in NAD(P)H:quinone oxidoreductase and in myeloperoxidase influenced susceptibility to benzene hematotoxicity. Other work related to toxicogenomic exploration of susceptibility to benzene hematotoxicity was described earlier in this review. These toxicogenomic studies should be extended to the risk of benzene-induced neoplasia, although the presumption is that hematotoxicity is related to risk of cancer. 13.4.1.2 Neoplastic Effects Benzene is a known cause of acute myelogenous leukemia (AML), the adult form of acute leukemia that was almost uniformly fatal until recent advances in chemotherapy. Individual cases of AML and its variants (grouped under the heading acute nonlymphocytic leukemia—ANLL) in benzene-exposed individuals were first reported about seven decades ago, but it was not until the 1970s that the causal relationship was fully accepted. Benzene exposure leads to cytogenetic abnormalities in bone marrow cells and in circulating lymphocytes (Forni et al., 1971; Zhang et al., 2002), consistent with alteration of the genome and in keeping with a somatic mutation leading to cancer. Cytogenetic changes observed in humans exposed to benzene fall into three categories: micronuclei, sister chromatid exchanges, and chromosomal aberrations. Zhang et al. (2002) reviewed the published findings and concluded that there is, thus far, unfortunately, no evidence of a unique pattern of benzene-induced chromosomal aberrations in humans. The known increased likelihood of a demonstrable cytogenetic abnormality in those cases of leukemia associated with a high level of workplace benzene exposure has been extended recently to cryptic cytogenetic abnormalities determinable in AML patients only with advanced techniques (Cuneo et al., 2002). More recent studies in benzene-exposed Chinese workers have begun to apply molecular biological techniques to better understand the chromosomal effects of benzene. For example, Rothman et al. (1995) noted that benzene induced gene duplicating but not gene inactivating mutations at the glycophorin A locus in bone marrow cells. Of note is that hematologists have long recognized that anyone with an aplastic anemia from virtually any cause has an increased risk of developing AML. Thus, radiation and alkylating agents used in chemotherapy produce both aplastic anemia and AML.


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Unequivocal evidence of the causal relationship between benzene and AML has come through studies in a number of countries. An epidemic of aplastic anemia followed by an epidemic of AML was observed in Turkish leather workers due to the introduction of a benzene-containing glue in their workshops (Aksoy, 1985; Aksoy et al., 1971). A cohort of workers in a Pliofilm factory in Ohio has become among the most thoroughly studied in the history of occupational medicine, particularly in relation to reconstructing their past exposure history (Infante et al., 1977; Rinsky et al., 1981, 1987; Ward et al., 1996). Ten cases of myelogenous leukemia were observed, with only two expected. Studies in Chinese workers have again demonstrated the leukomogenic potential of benzene (Yin et al., 1987a, 1996a; Hayes et al., 1997, 2000, 2001). Often individuals with aplastic anemia or some degree of pancytopenia are observed to go through a preleukemic stage of varying duration before developing frank AML. The manifestations observed, including morphological abnormalities in bone marrow precursors, have been classified together as the myelodysplastic syndrome (Layton and Mufti, 1987), a syndrome that is not unique to benzene hematotoxicity. Benzene also appears to share, with radiation and with chemotherapeutic alkylating agents, a predilection for a lag period between initial exposure and the development of frank myelogenous leukemia of 5–15 years (Goldstein and Kipen, 1991). This is relatively short for other types of cancers. Although the data for benzene are less clear cut, a case can be made that it would be distinctly unusual for a benzene exposure to result in AML in less than 2 years, and there does appear to be a lessening of risk perhaps 10–15 years following cessation of exposure. Benzene has also been highly associated with other neoplasms, both hematological and nonhematological (Young, 1989; Goldstein, 1977, 1990), although the evidence is not as definitive as it is for AML. An example is multiple myeloma. This is a bone marrow tumor of plasma cells that are antibody-producing cells derived from b-lymphocytes. A few individual cases of multiple myeloma have been reported in association with benzene exposure, including four cases in the Turkish group (Aksoy, 1985; Aksoy et al., 1984). Four deaths from multiple myeloma were initially reported in the Ohio Pliofilm cohort as compared to one expected, a statistically significant observation (Rinsky et al., 1987). A causal relationship is biomedically plausible in view of the fact that benzene unequivocally can cause a bone marrow tumor, and b-lymphocytes that are precursors of plasma cells develop cytogenetic abnormalities following benzene exposure. A related question is, if multiple myeloma is causally related to benzene exposure, why is the evidence less definitive than it is for AML? This important point is also pertinent when considering the relationship between benzene and other hematological and nonhematological neoplasms. There are three main considerations. The first is the relative risk per unit of benzene exposure for AML in comparison to multiple myeloma. It is conceivable, although speculative, that a given amount of benzene is more likely to produce the somatic mutation leading to AML than that it will produce the mutation responsible for multiple myeloma; or, similarly, benzene may be more likely to lead to promotion or progression of the mutation eventually resulting in AML. Collins et al. (2003) have also suggested that peak benzene exposures may be an important factor in causation of multiple myeloma and other lymphohematopoietic cancers. The second consideration, which is not speculative, is that multiple myeloma is less common a cancer than is AML. Thus, any multiplication of its original background risk caused by benzene exposure is less likely to be recognized given the usual size of cohorts (e.g., if the background incidence of death from tumor A is 10 in 1000 and of tumor B is 1 in 1000, then the doubling of risk by a carcinogen in a large cohort recording 1000 deaths would lead to 20 deaths as compared to 10 expected for tumor A, a

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statistically significant finding; however, for tumor B there would be a statistically unrecognizable two deaths as compared to one expected). The third point concerns the relatively short latency period for AML as compared to other tumors (Goldstein, 1989b, 1990). A greater delay between onset of exposure and eventual tumor complicates the likelihood of observing a causal relationship. An example of the common problem inherent in detecting a relation between benzene and hematological tumors other than AML is given by the recent study of mortality in 5514 workers occupationally to benzene in England and Wales (Sorahan et al., 2005). The standard mortality ratio for all leukemias was 137, which was not statistically significant, and 183 for ANLL, which was statistically significant at p < 0.05. For any of the three reasons described above, it is not surprising that this large and sophisticated study, which barely is able to detect an increase in ANLL, would not have the power to detect a causal effect of benzene on multiple myeloma or other hematological tumors. Similarly, Huebner et al. (2004), in a study of mortality in two petrochemical refineries, found that, in one, there was a statistically significant increase in the broad category of lymphohematopoietic cancers (SMR 1.47) and for leukemia overall (SMR 1.69), but the SMR of 1.52 for ANLL was not statistically significant and the SMR for non-Hodgkin’s lymphoma (1.47) was “near significant.” A study reputed to be a reason to discard the hypothesis of a role of benzene in causing multiple myeloma, that of Bergsagel et al. (1999), has an intrinsic flaw that makes it incapable of addressing the question. Basically, the cohort had such low benzene exposure that it would not be expected to have an increase in AML (or ANLL), the signature of benzene effect. Looking at this group for a tumor other than AML is bound to be negative, as would asking whether cigarette smoking caused multiple myeloma in a cohort whose pack-years of smoking was too low to cause a detectable increase in lung cancer (Goldstein and Shalit, 2000). The evidence supporting benzene as a cause of non-Hodgkin’s lymphoma (NHL) is even greater than it is for multiple myeloma, in part because of the observation of lymphomas in mice exposed to benzene (Snyder et al., 1980). The findings in China of at least a borderline statistically significant increase in NHL (relative risk 3.0, 99% confidence interval (CI): 0.9–10.5) is particularly important in view of the low background incidence of lymphatic tumors in China. For those with 10 or more years of benzene exposure, the relative risk for NHL was 4.2 (95% CI: 1.1–15.9) (Hayes et al., 1997). A related study in the same cohort showed a statistically significant increase in risk for NHL (Yin et al., 1996b). Perhaps the most perverse positive evidence of the role of benzene in causing NHL, as well as the difficulty posed by the healthy worker effect to interpreting worker cohort studies, is that of Wong and Raabe (2000). They reported that the SMR for NHL was 0.90 (95% CI: 0.82–0.98) in a pooled multinational cohort of over 300,000 workers, which they stated demonstrated that benzene was not a cause of NHL. As with their study of multiple myeloma (Bergsagel et al., 1999), there is a fatal flaw in relating the findings to benzene carcinogenicity in that their cohort did not have an overall increase in AML, the signature event of benzene exposure. Further, they did not consider the healthy worker effect, which is particularly prominent as NHL is a not uncommon outcome in those with HIV infection or with immune-related disorders—and such individuals are far less likely to become members of a petroleum refinery workforce. In fact, in their subanalysis, they report that the SMR for NHL in U.S. refinery workers was 0.96 (95% CI: 0.86–1.07) and non-U.S. refinery workers was 1.12 (95% CI: 0.90–1.37), while for gasoline distribution workers it was 0.64 (95% CI: 0.50–0.82). The lack of overlap in confidence intervals between the refinery workers who can be exposed to pure streams of benzene and the distribution workers who


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have far less opportunity for benzene exposure strongly indicates a statistically significant impact of working in petroleum refineries on increasing the incidence of NHL. This impact is obscured by comparison with the total population that includes a much higher incidence of those with a propensity to develop NHL than does the worker population. Acute lymphatic leukemia (ALL), which is the usual childhood form of leukemia, is also highly likely to be caused by benzene. In addition to the evidence cited above for other lymphatic neoplasms, there is some epidemiological evidence in workers. Further, the molecular evidence of a common bone marrow precursor for lymphatic and myelocytic cells is strongly supportive. Obviously, as children are rarely exposed to high levels of benzene in the workplace, epidemiological evidence will be difficult to obtain. Steffen et al. (2004) noted a statistically significant association between childhood leukemia and residence near a gasoline station or repair garage, but not between ALL and traffic density. Recently, Schnatter et al. (2005) have reviewed the literature on the relation of leukemia subtypes to benzene exposure. The possibility that benzene might be the cause of nonhematological neoplasms, such as lung or liver cancer, has been raised by animal studies in which it has been difficult to demonstrate hematological neoplasms. A variety of solid tumors have been observed in 2-year rodent studies performed by Maltoni et al. (1983) in Italy and by the U.S. National Toxicology Program (Huff, 1983). However, there is no convincing evidence of nonhematological tumors occurring in humans as a result of benzene exposure. In predicting whether these would occur, it is of conceptual importance to determine whether the proximal carcinogen is uniquely formed within the bone marrow or whether the carcinogenic metabolites are made primarily in the liver and then travel throughout the circulation, the bone marrow being particularly susceptible because of its special dynamics but as in radiation carcinogenesis, other organs also being at risk. 13.4.2

Nonhematological Effects

13.4.2.1 Central Nervous System Benzene has an odor threshold in the range of 4– 5 ppm. The acute central nervous system toxicity of benzene is similar to that of other alkyl benzenes and has much in common with the general anesthetic effects of lipophilic solvents. Acute central nervous system toxicity appears to be a direct effect of benzene, and not related to its metabolites. Symptoms following acute inhalation include drowsiness, lightheadedness, headache, delirium, vertigo, and narcosis leading to loss of consciousness. Benzene levels at which acute central nervous system effects become noticeable are at least above 100 ppm and are more likely to be in the few hundred or few thousand parts per million range. In view of the leukemogenicity of benzene, controlled human exposure studies of such effects are not appropriate. On structure–activity grounds, benzene might be expected to have acute central nervous system effects similar to toluene, albeit at a slightly lower dose for benzene. Chronic nervous system effects of benzene have not been unequivocally demonstrated. There is much debate about whether chronic nervous system effects may occur with alkyl benzenes, particularly in work groups such as commercial painters, and, if proven, such findings might conceivably be pertinent to chronic benzene exposure. However, the exposure of painters to benzene is now at a far lower level than in the past—the predominant solvent exposures being to toluene and other alkyl benzenes. 13.4.2.2 Reproductive and Developmental Effects Keller and Snyder (1986, 1988) have shown hematological effects in the developing fetus at relatively low maternal exposure

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concentration. However, to date, animal studies have generally been negative, and there is no human evidence for reproductive or developmental toxicity from benzene despite substantial exposure of women of reproductive age. Studies evaluating the possible link of childhood leukemia to parental occupation have shown no clear pattern (Ali et al., 2004). 13.4.2.3 Effects on the Immune System There is no question that high levels of benzene produce effects on lymphocytes in laboratory animals exposed to benzene, and that these lymphocytopenic effects can result in deficits in immune function (Irons 1985). Immune function decrements can occur in the absence of lymphocytopenic effects (Rosenthal and Snyder, 1985, 1987). However, there is currently no evidence whatsoever to suggest that immune function is affected following exposure of humans to allowable workplace levels of benzene or to the much lower levels present in the general environment.

13.5 RISK ASSESSMENT Risk assessment for benzene remains controversial. As compared to many other carcinogens, there has long been comparatively good agreement about the risk-specific dose of benzene (Goldstein, 1985), yet the economic and societal stakes are so high as to make relatively small differences in the interpretation of great significance to industrial stakeholders. Zeise and McDonald (2000) reviewed some of the risk assessment issues considered by the state of California, and the use of nontumor data for the cancer risk assessment of benzene has been reviewed by Albertini et al. (2003). The four major components of risk assessment are hazard identification, dose–response estimation, exposure assessment, and risk characterization. Particularly controversial has been the exposure assessment for worker cohorts in which there has been epidemiological identification of an increase in leukemia risk. This will continue, particularly in relation to exposure of Chinese workers. In addition, the possibility of a hyperbolic shape of the dose–response curve will be debated, as will the implications of toxicogenomic data relevant to human susceptibility to benzene. 13.5.1

Hazard Identification

In terms of hazard identification, there is no question that benzene is a known cause of hematological cancers, particularly acute myelogenous leukemia, and of bone marrow damage leading to pancytopenia and aplastic anemia. Benzene unequivocally causes AML and its variants, grouped under the heading of ANLL. As discussed above, hazard identification of benzene as causal for certain other hematological cancers varies from likely to almost certain. High levels of benzene, like other hydrocarbon solvents, can produce acute central nervous system toxicity and are a legitimate workplace concern—particularly in enclosed workspaces. Other effects, such as chronic CNS toxicity and fetal abnormalities, are of theoretical concern, but there is no direct evidence to support these hazards in humans. 13.5.2

Dose–Response Estimation

13.5.2.1 Carcinogenicity of Benzene Dose–response estimation for benzene, as for other carcinogens, has generated controversy. Points of contention include the use of the standard EPA linearized multistage model, which in essence assumes that every

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single molecule of benzene has a finite chance of producing the somatic mutation responsible for acute myelogenous leukemia. Industry at times has argued that aplastic anemia is a necessary precursor of AML, based on observations of individuals with aplastic anemia who evolved through a preleukemic phase to AML, on the argument that many of the reported cases of benzene-associated AML have occurred in individuals with relatively high levels of exposure, and on the knowledge that aplastic anemia from seemingly any cause seems to predispose toward AML. There is clearly a threshold for aplastic anemia; that is, there is a level of benzene exposure below which there is no apparent risk of aplastic anemia. Accordingly, if aplastic anemia is a necessary precursor of AML, then there must also be a threshold for benzene leukemogenesis, and this threshold must be well beyond any exposure level of environmental concern. However, the supporting evidence for such a threshold for AML is very weak. In terms of case reports, there have been numerous cases of benzene-associated AML in which there was no evidence of antecedent aplastic anemia, although some degree of pancytopenia cannot be ruled out because of lack of preceding blood counts. In addition, a mechanism that depended solely on some leukemogenic process occurring in response to aplastic anemia would have difficulty explaining the cytogenic abnormalities that occur at levels of benzene that do not produce a frank pancytopenia (Yardley-Jones et al., 1988; Smith, 1996). The slope of the dose–response curve for benzene leukemogenesis undoubtedly depends on the metabolism of benzene to proximal carcinogen(s). As described in detail above, the complex metabolic pathways leading to active species are in the process of being unraveled. Such information should be of great value in establishing the appropriate dose–response relationship for benzene leukemogenesis. Although it is difficult to overstate the importance of understanding benzene metabolism in relation to benzene leukemogenesis, it must also be emphasized that determining the kinetic relationship between benzene exposure and the formation of active species is not sufficient by itself to establish the dose–response relationship. One must also take into account the responsiveness of the target organ, which, through its own natural variation in sensitivity or in defense mechanisms, may play a role in determining the dose–response pattern. For example, Sabourin et al. (1990) and Witz et al. (1990) have shown that the percentage of benzene metabolized to the ring-opened form, as measured by muconic acid, is inversely proportional to the exposure level. This suggests that the dose pattern of concern might be chronic low-level exposure rather than the equivalent dose given as a short-term spike, and that there may be a hyperbolic dose–response curve (i.e., per unit dose there is greater potency at the lower end of the dose–response curve). Recent studies by Rappaport et al. (2005) lend credence to this possibility. However, Witz et al. (1985) have also shown that the bone marrow toxicity of trans,trans-muconaldehyde is greater with a single daily dose than it is with the same dose divided into three daily injections; that is, the short-term spike of the toxic metabolite is more harmful to the target organ. Integrating physiologically based pharmacokinetics for benzene metabolites with the dose pattern responsiveness of the bone marrow will present an intriguing challenge to those interested in modeling dose– response patterns. The dose–response estimation for benzene leukemogenesis using currently standard EPA models has been derived from the leukemia incidence in studies of occupationally exposed cohorts. The study that finally convinced regulators that benzene was a human leukemogen, that of Infante et al. (1977) on Pliofilm workers, was a major basis for the original EPA dose–response estimation of the leukemogenic potential of benzene. A highly controversial retrospective analysis of dose was required (see below). More recently epidemiological

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studies of worker cohorts have generally paid more attention to the dose side of the dose– response analysis. Three groups of studies have been particularly notable for reporting an increase in leukemia risk at lower levels of benzene than previously reported or expected: studies in China by a group that includes the Chinese Academy of Preventive Medicine; the U.S. National Cancer Institute and academics at a variety of U.S. institutions, most notably the University of California at Berkeley School of Public Health and New York University; the second is from a group in France investigating workers at their electric power industry, and the third from a group evaluating workers at the Australian petrochemical industry. The latter two studies are relatively easy to discount in terms of their having any impact on the dose–response analysis for benzene leukemogenesis. However, the multiple ongoing studies in China will certainly have an impact on considerations of benzene cancer risk. The studies in China have been reviewed by Hayes et al. (2000, 2001). The rapid industrialization of China has taken a toll in worker health, and the adverse health impact of the inappropriate use of benzene is just one example. Not unexpectedly, an increase in hematological tumors in benzene-exposed Chinese workers has been documented (Yin et al., 1987a, 1996a; Hayes et al., 1997, 2000, 2001). However, in terms of the risk assessment of benzene, evaluation of the benzene exposure of these Chinese workers led to the observation of a higher relative risk per unit benzene than anticipated based upon the Pliofilm and other cohorts of benzene-exposed workers. Questions were immediately raised about whether there had been an underestimate of actual exposures. As described below (see Section 13.5.3), this has led to a more thorough evaluation of the benzene-exposed workforce in China with residual controversy. Among the surprising findings of leukemia seemingly associated with low-level benzene exposure is that of Guenel et al. (2002) funded by the power industry in France, which heavily relies on nuclear power. An increase in leukemia risk in a cohort of workers in this industry was tentatively ascribed by the authors to benzene exposure rather than to the more likely exposure to radiation, a known leukemogen. The authors estimated that mean benzene TWA exposure for exposed workers was 0.16 ppm, but this estimate was made despite an apparently complete lack of any actual benzene measurements. As pointed out by the authors, the observed risk may be related to other occupational factors. A particularly thorough but partially flawed nested case–control study of incidence of disease in refinery workers has been performed in Australia by Glass et al. (2003). While again demonstrating an increased risk for AML in those refinery workers most heavily exposed to benzene, the study is marred by a likely spurious finding due to surveillance bias of an increase in chronic lymphatic leukemia (CLL) in this workforce at particularly low levels of benzene (Goldstein, 2004). In contrast to AML, which almost always has a rapid onset and progression, CLL is a disease that develops relatively slowly. Not uncommonly, diagnosis of CLL results from an unexpected finding of lymphocytosis in a routine blood count in an otherwise healthy individual. In such cases, average life expectancy is more than 12 years. Obviously, the more frequently routine blood counts are performed, the more likely the diagnosis will be made. Glass et al. (2003) studied the entire refinery workforce, including those with negligible benzene exposure (e.g., office staff). Refinery workers exposed to benzene routinely have blood counts as part of standard workplace surveillance. Accordingly, those with CLL in the benzene-exposed part of the workforce would have been discovered relatively early in their long disease process, while those not subjected to hematological surveillance could have an undiagnosed case of early CLL that would not be detected. Because of the high likelihood of a surveillance bias, this study cannot be used as evidence that benzene causes CLL, nor can the CLL cases in the study be incorporated into

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any extrapolation used for risk assessment purposes. The latter point is particularly important as the authors report an excess total leukemia incidence in association with cumulative lifetime benzene exposure levels as low as >2–4 ppm-years. However, in their conditional logistic regression analysis of the association of leukemia subtype with the lowest level of cumulative benzene exposure in this particular analysis, >4–8 ppm-years, the OR was only 0.52 (0.05–5.0) for ANLL. In contrast, at these low benzene levels the OR for CLL was 2.76 (0.42–18.1), again in keeping with surveillance bias accounting for the findings. 13.5.2.2 Noncarcinogenic End Points There seems to be no reasonable likelihood that aplastic anemia can be caused at the levels of exposure reported in the general environment. The threshold for aplastic anemia as a potentially crippling or fatal disease, even for benzenesensitive individuals, is likely to be significantly above the 10 ppm TWA (time-weighted average) benzene standard that was the U.S. workplace norm for many years. As discussed above, the recent report of observable differences in blood counts in Chinese workers exposed to less than 1 ppm benzene (Lan et al., 2004) will increase interest in estimating the risk for the noncancer effects of low levels of benzene. There are three issues that need to be addressed: the need for replication of these hematological findings in other closely observed worker groups, including resolving the discrepancy between the Lan et al. (2004) and Quitt et al. (2004) findings relating to precursor cells (see above); an understanding of the extent to which a small decrement of a blood count well within the normal range would be considered an adverse effect for risk assessment purposes; and the appropriate use of uncertainty (safety) factors for such an end point. The use of large safety factors resulting in a sub-ppm benzene standard to protect against noncancer hematological effects is open to question. Starting with animal data and then using multiple safety factors might be relevant for a new chemical, but is simply inappropriate given the literally millions of person-years of data available for benzene. An example of the misapplication of this approach occurred a few years ago in New Jersey, where the use of a 10 ppb indoor “acute action level” in a situation with gasoline-contaminated groundwater led to evacuation of homeowners and widespread community concern. This standard was based upon use of a 100-fold safety factor on a 1 ppm standard. The families were moved to a crossroads motel that likely contained higher indoor benzene levels than their homes. The level of 10 ppb is not uncommon in American basements. In view of the lack of positive data, no risk levels for fetal abnormalities or chronic nervous system effects from benzene are currently indicated. 13.5.3

Exposure Assessment

Exposure assessment for benzene presents some current areas of intense controversy, both in terms of retrospective estimates of exposure of benzene-exposed cohorts with elevated leukemia incidence and in terms of the appropriate approach to exposure estimation from known benzene sources in a community. Perhaps the most thorough retrospective evaluation of benzene exposure, or of any occupationally exposed cohort, has been that performed by Rinsky et al. (1987) on the employees of a Pliofilm plant whose benzene exposure resulted in a substantial risk of leukemia. The authors did a masterful job of reconstructing the location of each of the workers within the workplace and possible exposure levels. However, Rinsky et al. (1987) ended up with much lower exposure levels, particularly during the Second World War,

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when most of the leukemia cases were at work, than did Crump and Allen (1984), who built their exposure estimate on the basis of the Rinsky et al. efforts. The major distinction between the two approaches is the much higher levels posited by Crump and Allen during World War II. The latter appears to be strongly supported by contemporary accounts of significant incidences of aplastic anemia and of central nervous system effects at Pliofilm factories as a result of wartime conditions, which would also expect to result in longer working hours and thus greater individual exposures. In addition, Kipen et al. (1988, 1989a) found that the historic blood counts of workers in this cohort had a statistically significant inverse correlation with the Crump and Allen exposure assessment but no relationship with the estimate of Rinsky et al. (see also Hornung et al., 1989; Kipen et al., 1989b; Paxton et al., 1994a, 1994b; Schnatter et al., 1996). Because almost all of the workers with AML had been in this cohort in the early 1940s, it appears that reliance on the Rinsky et al. exposure estimation to perform a risk assessment may lead to an overestimate of the risk. Because of its relevance to risk assessment, the reassessment of the exposure of this cohort continues (Williams and Paustenbach, 2003). Similar to the Pliofilm cohort, it can be anticipated that the extent to which the recent studies of benzene-exposed workers in China will impact on the cancer risk potency for benzene will depend in large part upon the scientific acceptability of the exposure assessments (e.g., Dosemeci et al., 1994; Vermeulen et al., 2004; Waidyanatha et al., 2004), which have been questioned (Wong, 2002). 13.5.4

Risk Characterization

The EPA has characterized the risk of benzene-induced leukemia as being 8 106 (mg/m3 benzene) (ppb) per 70-year lifetime. This can be translated into a risk of 8 in 1 million of dying from benzene-induced leukemia from breathing 1 mg/m3 benzene continually for 70 years. This risk is usually described as being the plausible upper boundary in that the conservative assumptions built into risk assessment are thought to make it unlikely that the risk is any higher, but it is conceivable that it is much lower. The level of benzene responsible for a one in 1 million lifetime risk, for drinking water, has been similarly calculated by EPA to be 0.66 ppb.

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Yin S-N, Hayes RB, Linet MS, Li G-L, Dosemeci M, Travis LB, Li C-Y, Zhang Z-N, Li D-G, Chow W-H, Wacholder S, Wang Y-Z, Jiang Z-L, Dai T-R, Zhang W-Y, Chao X-J, Ye P-Z, Kou Q-R, Zhang X-C, Lin X-F, Meng J-F, Ding C-Y, Zho J-S, Blot WJ (1996) A cohort study of cancer among benzene-exposed workers in China: overall results. Am. J. Ind. Med. 29:227–235. Yin SN, Hayes RB, Linet MS, Li GL, Dosemeci M, Travis LB, Zhang ZN, Li DG, Chow WH, Wacholder S, Blot WJ (1996b) An expanded cohort study of cancer among benzene-exposed workers in China. Benzene Study Group. Environ Health. Perspect. 104S6:1336–1341. Yoon B, Li G, Kitada K, Kawasaki Y, Igarashi K, Kodama Y, Inoue T, Kobayashi K, Kanno J, Kim D, Inoue T, Hirabayashi Y (2003) Mechanisms of benzene-induced hematoxicity and leukemogenicity: cDNA microarray analyses using moue bone marrow tissue. Environ. Health Perspect. 111:1411–1420. Young N (1989) Benzene and lymphoma. Am. J. Ind. Med. 15:495–498. Yu R, Weisel C (1996) Measurement of benzene in human breath associated with an environmental exposure. J. Expo. Anal. Environ. Epidemiol. 6:261–277. Zeise L, McDonald TA (2000) California perspective on the assessment of benzene toxicological risks. J. Toxicol. Environ. Health A 61:479–483. Zhang Z, Kline SA, Kirley TA, Goldstein BD, Witz G (1993) Pathways of trans,trans-muconaldehyde metabolism in mouse liver cytosol: reversibility of monoreductive metabolism and formation of end products. Arch. Toxicol. 67:461–467. Zhang Z, Xiang Q, Glatt H, Platt KI, Goldstein BD, Witz G (1995a) Studies on pathways of ring opening of benzene in a Fenton system. Free Radic. Biol. Med. 18:411–419. Zhang Z, Goldstein BD, Witz G (1995b) Iron-stimulated ring-opening of benzene in a mouse liver microsomal system. Mechanistic studies and formation of a new metabolite. Biochem. Pharmacol. 50:1607–1617. Zhang Z, Schafer F, Schoenfeld H, Cooper K, Snyder R, Goldstein BD, Witz G (1995c) The hematotoxic effects of 6-hydroxy-trans,trans-2,4-hexadienal, a reactive metabolite of trans, trans-muconaldehyde, in CD-1 mice. Toxicol. Appl. Pharmacol. 132:213–219. Zhang L, Eastmond DA, Smyth MT (2002) The nature of chromosomal aberrations detected in humans exposed to benzene. Crit. Rev. Toxicol. 32:1–42.

14 CARBON MONOXIDE Michael T. Kleinman

14.1 INTRODUCTION Carbon monoxide is emitted from virtually all sources of incomplete combustion, including internal combustion engines (e.g., automobiles, trucks, and gasoline-fueled small engines) (Duci et al., 2003; El-Fadel and El-Hougeiri, 2003; Mott et al., 2002; Ott et al., 1994; Renner, 1988; Utell et al., 1994); fires, both natural and man-made; improperly adjusted gas and oil appliances (e.g., space heaters (Anonymous, 1997; Setiani, 1994), water heaters (Howell et al., 1997), stoves (Guggisberg et al., 2003; Samet et al., 1987), and ovens (Anonymous, 1997; Angle, 1988)); and tobacco smoking (Calafat et al., 2004; Gourgoulianis et al., 2002; Murray et al., 2002; Viegi et al., 2004). Because of the large number and the ubiquity of CO sources with significant source strengths (e.g., tobacco smoke contains about 1% CO by volume, or 10,000 ppm CO), ambient CO concentrations show large temporal and spatial variations. The exposure of individuals to CO is, therefore, also quite variable, depending upon the types of activities in which a person is engaged and how long they are engaged in those activities (time–activity profiles), where the activity takes place (microenvironments, for example, indoors, at a shopping mall, outdoors, in a vehicle, at work or school, in a parking garage, or even in a skating rink) (Horner, 2000; Jovanovic et al., 1999; Levesque et al., 2000, 2005; Viala, 1994), and the proximity to CO sources (Campbell et al., 2005; Linn and Gong, 1999; Ott, 1990; Ott et al., 1992). Various methods have been used to document exposures, including the use of data from fixed-site ambient monitors, the use of microenvironmental exposure assessment models, personal exposure monitoring methods, and biological monitoring methods (Ott et al., 1992). Controls on motor vehicle exhaust and the use of catalytic converters on vehicles sold in the United States have been very effective in reducing ambient CO emissions and commuter exposures (Hinkle, 1980; Hutchinson and Pearson, 2004; Hysell et al., 1975; Mott et al., 2002). However, while the increased use of oxidizing catalysts resulted in decreased emissions of CO, they also increased emissions of noble metals and ultrafine particles, which may contribute to health effects in other ways


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(Brubaker et al., 1975; Finklea et al., 1975). The fact remains that the reduction in automotive emissions brought about by the Clean Air Act has reduced traffic CO exposures and traffic-related CO concentrations well below those measured prior to the year 2000 (U.S. EPA, 1999).

14.2 CO EXPOSURE AND DOSIMETRY CO competes with oxygen for binding sites on the heme portion of the hemoglobin (Hb) molecules in red blood cells to form carboxyhemoglobin (COHb). The affinity of Hb for CO is about 240–250 times that for O2 (Roughton, 1970). The formation of COHb by the binding of CO to circulating Hb reduces the oxygen-carrying capacity of blood. In addition, binding of CO to one of the four hemoglobin binding sites increases the O2 affinity of the remaining binding sites, thus interfering with the release of O2 at the tissue level. When O2 content of blood (mL O2/mL blood) is plotted versus O2 partial pressure (mmHg) in blood, the increased O2 affinity is seen as the so-called leftward shift in the curve for blood partially loaded with CO (Okada et al., 1976; Zwart et al., 1984). CO-induced tissue hypoxia is therefore a joint effect of the reduction in O2-carrying capacity and the reduction of O2 release at the tissue level. The brain and heart, under normal conditions, utilize larger fractions of the arterially delivered O2 (about 75%) than do peripheral tissues and other organs (Ayres et al., 1970) and are therefore the most sensitive targets for hypoxic effects following CO exposures. The potential for adverse health effects is increased under conditions of stress, such as exercise, which increases O2 demands at the tissue level to sustain metabolism. The measure of biological dose that relates best to observed biological responses and deleterious health effects is the concentration of COHb expressed as a percentage of available, active Hb, thus representing the percent of potential saturation of Hb. COHb can be measured directly in blood or estimated from the CO content of expired breath (Berny et al., 2002; Attebring et al., 2001; Groman et al., 1998; Laranjeira et al., 2000; Rea et al., 1973; Vogt et al., 1977; Vreman et al., 1996; Wickramatillake, 1999). It is currently accepted that the most accurate and reliable method for measuring COHb concentration is by gas chromatographic analysis (Berny et al., 2002; Vreman et al., 1984, 1994). Spectrophotometric measures and instruments have been widely used in both clinical and occupational health settings (Boumba and Vougiouklakis, 2005). However, spectrophotometric instruments may have limited accuracy and precision at COHb concentrations below 5% (Allred et al., 1989; Chaitman et al., 1992; Johansson and Wollmer, 1989). If used for research studies, they should be calibrated properly and measurements should verified by a “gold standard,” such as gas chromatography (Johansson and Wollmer, 1989; Widdop, 2002; Wigfield et al., 1981). When direct measurements cannot be made, it is possible to estimate COHb from ambient air CO concentrations (Ott et al., 1988), indoor air CO concentrations, and personal CO monitoring data (Ott and Mage, 1978, 1979) using pharmacokinetic and other models (Benignus et al., 1994; Bruce and Bruce, 2003; Chung, 1988; Coburn et al., 1965; Hauck and Neuberger, 1984; Joumard et al., 1981; Peterson and Stewart, 1975; Tikuisis et al., 1987) that link the concentration of inhaled CO, breathing rate and volume, blood volume, metabolic production of endogenous CO, and rate of removal of CO. The Coburn– Forster–Kane (CFK) model (Coburn et al., 1965) has been widely used for this purpose. But perhaps more importantly, in the process of establishing ambient air quality guidelines, the CFK model has been the basis for associating ambient and workplace air CO concentrations

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with concentrations of COHb that could be hazardous for sensitive exposed individuals. The CFK model has been experimentally verified for exposures at 25–5000 ppm, during rest and exercise (Peterson and Stewart, 1975; Tikuisis et al., 1987). Sensitivity analyses of CFK can be used to identify those model input parameters for which errors in the parameter values that are used will have the greatest errors in predicted COHb concentrations (McCartney, 1990). Validation of the CFK model under well-controlled conditions at environmental concentrations with larger numbers of subjects would be useful. However, Kleinman et al. (1998) used the CFK model to successfully predict blood COHb concentrations in 17 human volunteers exposed to 100 ppm CO, suggesting that the CFK model is valid under such circumstances.

14.3 MECHANISMS OF CO TOXICITY CO affects health indirectly by interfering with the transport of oxygen to tissues (especially the heart and other muscles and brain tissue) (McGrath, 2000). The resulting impairment of O2 delivery cause tissue hypoxia and interferes with cellular respiration. When CO is taken up by cells, it can complex with Fe2þ in hemoproteins such as myoglobin (McGrath, 2000), cytochrome oxidase, and cytochrome P450 (so named because the Fe2þ–CO complex absorbs light with a maximum absorption at 450 nm) (Williams, 1992), and thus interfere with electron transport processes and energy production at the cellular level. Thus, in addition to observed physiological effects and cardiovascular effects, CO can modify electron transport in nerve cells resulting in behavioral, neurological, and developmental toxicological consequences. The possible role of CO as an etiologic factor in development of atherosclerosis is suggested by effects of tobacco smoke exposure (Hart, 1993; Leone, 1995) and mobile source emissions (Utell et al., 1994), but long-term exposures to 200 ppm CO in a sensitive animal model failed to show an effect of CO (Penn, 1993). The role of CO as a causative factor in cardiac arrhythmias, sudden cardiac arrest, and myocardial infarctions is an area of active research activity. The hemodynamic responses to CO have been reviewed for both animal models and humans (Kanten et al., 1983; Penney, 1988). Chronic CO exposures, usually at COHb concentrations greater than 10%, produce several changes. These may be adaptive responses to induced hypoxia, such as increases in numbers of red blood cells (polycythemia), increased blood volume, and increased heart size (cardiomegaly). In addition, heart rate, stroke volume, and systolic blood pressure may be increased. Some of these effects have been seen in smokers. Other environmental factors, such as effects of other pollutants (both from conventional air pollution sources and from environmental tobacco smoke), interactions with drugs and medications, health and related factors (e.g., cardiovascular and respiratory diseases, anemia, or pregnancy), and exposures at high altitude are possible risk modifiers for the health effects of CO. Exposures to high concentrations of CO due, for example, to fires and emissions from faulty appliances result in over 2000 deaths per year in the United States and other countries (Abu-al Ragheb and Battah, 1999; Raub et al., 2000b; Sadovnikoff et al., 1992) and in illness sufficient to cause upward of 10,000 individuals to seek medical attention or to miss one or more days of work in the United States (Centers for Disease Control, 1982). The available data may substantially underestimate the total number of such cases, especially those related to unsuspected CO exposure in the home because some CO-related symptoms are similar to those of flu (headache and dizziness) and possible to those of certain seizure disorders (Heckerling, 1987; Heckerling et al., 1990a, 1990b; Kirkpatrick, 1987; Leikin et al., 1988).

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Therefore, many cases may be misdiagnosed and missed as being related to CO exposure. As a consequence, patients may not receive the proper treatment and their cohabitants may go untreated if they did not independently seek medical help (Heckerling et al., 1990b; Kao and Nanagas, 2004). Many of the inadvertent incidents of morbidity and mortality are preventable through the use of CO detectors, which are now readily available at a moderate cost (Yoon et al., 1998). Blood tests for COHb concentrations or breath analyses for CO improve the accuracy of the diagnoses. While many improvements have been made with respect to exposure to ambient CO, this remains as an important issue. The remainder of this chapter will, therefore, deal with ambient environmental exposures and will focus on the recent findings of CO-related health effects.

14.4 POPULATIONS AT RISK OF HEALTH EFFECTS DUE TO CO EXPOSURE 14.4.1

People with Cardiovascular Diseases

Daily variations in CO were strongly associated with hospital admissions among persons with ischemic heart disease (IHD) conditions, even after controlling for potential effects of ozone, nitrogen dioxide, or particulate matter less than or equal to 10 mm in aerodynamic diameter (PM10). A 1 ppm increase in 8 h average CO was associated with a 3.60% increase in same-day IHD admissions (Mann et al., 2002). Ischemic heart disease, also categorized as coronary artery disease, is a leading cause of disability and death in industrialized nations and may be associated with chronic elevation of COHb (Mall et al., 1985). It is a clinical disorder of the heart resulting from an imbalance between oxygen demand of myocardial tissue and oxygen delivery via the bloodstream. The ability of the heart to adjust to increases in myocardial O2 demands resulting from increased activity, or to reductions in O2 delivery by arterial blood due, for example, to COHb or reduced partial pressure in O2 in inspired air, by increasing O2 extraction, is limited, because the extraction rate in myocardial tissue is already high. Normally, coronary circulation responds to increased O2 demands by increasing blood flow. In coronary artery disease, the coronary artery is occluded by lipid deposits, which can impede augmentation of local coronary blood flow in response to increased O2 demands. Under these conditions, the myocardium is forced to extract more O2 resulting in reduced coronary venous and tissue O2 tensions, which can produce myocardial ischemia. Severe myocardial ischemia can induce a myocardial infarction (heart attack) or can alter cardiac rhythms, that is, cause arrhythmias. The association of acute CO exposure to heart attacks has been described (Koskela et al., 2000; Marius-Nunez, 1990; Martys, 1994; Scharf et al., 1974; Tan et al., 1993). Individuals with obstructed peripheral arteries may experience intermittent claudication, which is severe pain, usually in their legs, during walking or other relatively mild activities. CO exposure, for example, from cigarette smoking, can exacerbate the imbalance between O2 demand by exercising peripheral muscular tissue and O2 delivery in individuals with diseased peripheral arteries (Wald et al., 1977). 14.4.2

People with Anemia and Other Blood Disorders

Individuals with reduced blood hemoglobin concentrations, or with abnormal hemoglobin, will have reduced O2-carrying capacity in blood. In addition, disease processes that result in increased destruction of red blood cells (hemolysis) and accelerated breakdown of

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hemoproteins accelerate endogenous production of CO (Sannolo et al., 1992; Sears et al., 2001; Solanki et al., 1988), resulting in higher COHb concentrations than in normal individuals. For example, patients with hemolytic anemia have COHb concentrations two to three times those seen in normal individuals (Coburn et al., 1966). Endogenously produced CO, from the breakdown of hemoglobin by hemeoxygenase, was originally thought to be a superfluous by-product of heme catabolism. However, CO is now known to play a central role in blood pressure regulation, maintenance of organ-specific vascular tone, neurotransmission, stress response, platelet activation, and smooth muscle relaxation (Morse and Sethi, 2002; Wu and Wang, 2005). Thus, CO may be an important and beneficial mediator at normal physiological levels, but is toxic at elevated levels. 14.4.3

People with Chronic Lung Disease

Chronic lung diseases such as chronic bronchitis, emphysema, and chronic obstructive pulmonary disease (COPD) are characterized by impairment of the lung’s ability to transfer O2 to the bloodstream because diseased regions of the lung are poorly ventilated and blood circulating through these regions will therefore receive less O2 and accumulate carbon dioxide (so-called ventilation–perfusion mismatch) (Wagner et al., 1977; West, 1971, 1978). Exertional stress often produces a perception of difficulty in breathing, or breathlessness (dyspnea) in these individuals. Although exercise, and the metabolic acidosis associated with exercise in COPD patients, increases ventilatory drive, they have limited ventilatory capacity with which to respond (Sue et al., 1988). Reduction of blood O2 delivery capacity due to formation of COHb could exacerbate symptoms and further reduce exercise tolerance in these individuals. 14.4.4

Potential Risks for Pregnant Women, Fetuses, and Newborn Children

A CO-induced leftward shift in the O2Hb saturation curve may be significant for fetuses because the O2 tension in their arterial blood is low (20–30 mmHg) compared to adult values (100 mmHg) and because fetal Hb has a higher O2 affinity than does maternal Hb (Longo, 1976, 1977). Fetal blood has higher Hb concentrations than does maternal blood (Heilmann et al., 2005), which may compensate for the higher O2 affinity to some extent. In pregnant women, O2 consumption is increased 15–25%, and hemoglobin concentration may be simultaneously reduced, lowering the O2-carrying capacity of their blood (Sady and Carpenter, 1989). Epidemiological studies show that odds ratios for cardiac ventricular septal defects increased in a dose-responsive fashion with increasing CO exposure, a 1 ppm increase in mean exposure to CO during the first trimester of pregnancy is associated with a reduction of 23 g in birth weight, and first-trimester CO exposures were associated with 20% increased risk of intrauterine growth retardation (Gouveia et al., 2004; Newill, 1974; Ritz et al., 2002; Salam et al., 2005).

14.5 REGULATORY BACKGROUND The National Ambient Air Quality Standards (NAAQS) for CO were promulgated by the Environmental Protection Agency (EPA) in 1971 at levels of 9 ppm (10 mg/m3) for an 8 h average and 35 ppm (40 mg/m3) for a 1 h average, not to be exceeded more than once per year. (Primary and secondary standards were established at identical levels.) The 1970 CO criteria

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document (NAPCA, 1970) cited as the standard’s scientific basis a study that indicated that subjects exposed to low levels of CO, resulting in COHb concentrations of 2–3% of saturation, exhibited neurobehavioral effects (Beard and Wertheim, 1967). A reexamination of the scientific evidence, as reported in a revised CO criteria document (U.S. EPA, 1979), concluded that it was unlikely that significant, and repeatable, neurobehavioral effects occurred at COHb concentrations below 5%. Medical evidence, accumulated during the intervening years, however, indicated that aggravation of angina pectoris, and other symptoms of myocardial ischemia, occurred in men with chronic cardiovascular disease, exposed to low levels of CO resulting in COHb concentrations of about 2.7% (Anderson et al., 1973; Aronow et al., 1972; Aronow and Isbell, 1973; Aronow, 1974, 1979, 1981; Goldsmith and Aronow, 1975). EPA proposed, in 1980, based in part on the above studies, to retain the 8 h 9 ppm primary standard level, to reduce the 1 h primary standard from 35 to 25 ppm, and to revoke the secondary CO standards (because no adverse welfare effects had been reported at near-ambient levels). An EPA investigation found flaws in some of the Aronow studies from which data were used as part of the basis for the proposed reduction in the 1 h standard (Budiansky, 1983); EPA later decided to keep the 1 h standard at 35 ppm. In l984, EPA published an addendum to the 1979 CO criteria document that reevaluated the CO health effects data previously reviewed and took into account research that had been published in the interim (U.S. EPA, 1984). The document reviewed four effects associated with low-level CO exposure: cardiovascular, neurobehavioral, fibrinolytic, and perinatal. Dose–response data provided by controlled human studies allowed the following conclusions to be drawn: (a) Cardiovascular Effects. Among those with chronic cardiovascular disease, a shortening of time to onset of angina was observed at COHb concentrations of 2.9–4.5%. A decrement in maximum aerobic capacity was observed in healthy adults at COHb concentrations at and above 5%. Patients with chronic lung disease demonstrated a decrease in walking distance when COHb concentrations were increased from 1.1– 5.4% to 9.6–14.9%. (b) Neurobehavioral Effects. Decrements in vigilance, visual perception, manual dexterity, and performance of complex sensorimotor tasks were observed at, and above, 5% COHb. (c) Effects on Fibrinolysis. Although evidence existed linking CO exposure to fibrinolytic mechanisms, controlled human studies did not demonstrate consistent effects of carbon monoxide exposure on coagulation parameters. (d) Perinatal Effects. While there were some epidemiological associations between CO exposure and perinatal effects, such as low birth weight, slowed postnatal development, and incidences of sudden infant death syndrome (SIDS), the available data were not sufficient to establish causal relationships. In September 1985, EPA issued a final notice that announced the retention of the existing 8 h 9 ppm and 1 h 35 ppm primary NAAQS for CO and the rescinding of the secondary NAAQS for CO. The EPA reviewed health-related data in 1991 and completed the most recent CO criteria document in 1999 (U.S. EPA, 1999). In that interval, several controlled human clinical exposures, population-based studies, and inhalation studies using laboratory animal models were added to the available database. These studies have provided important insights into the

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possible mechanisms of toxic action of CO, in addition to those related to hypoxia, and illuminate effects not currently identified in human studies, or that might not be amenable to controlled human experimentation, such as perinatal and developmental effects. Following review of the 1991 and the 1999 CDs, the existing NAAQS for CO were retained and are the current standards.

14.6 HEALTH EFFECTS OF CO 14.6.1

Population-Based Studies

14.6.1.1 Acute Exposures and Their Effects Most of the population-based studies in the literature relating to the health effects of CO in humans have been concerned with exposures to combustion and pyrolysis products from sources such as tobacco, fires, motor vehicle exhaust, home appliances fueled with wood (Ellegard, 1996; Pierson et al., 1989), gas, or kerosene (Amitai et al., 1998; Cooper and Alberti, 1984), and small engines (Baldauf et al., 2006). The individuals in these studies are therefore exposed to variable, and usually unmeasured, concentrations of CO and also to high concentrations of other combustion products. Exposures to CO in occupational settings represent another substantial exposure classification, but such exposures are also often accompanied by exposures to other contaminants as well. The symptoms of CO poisoning are often nonspecific, or masked by an exacerbation of an underlying illness, such as congestive heart failure. The effects can range from mild, annoying symptoms that resolve after removal of the source to severe morbidity with profound central nervous system dysfunction and acute complications. Acute CO intoxication often results in neurologic and/or myocardial injury. Studies have reported that 2% to approximately 10% of patients display delayed neurological sequelae (Choi, 1983; Mathieu et al., 1985; Raub et al., 2000b; Thom and Keim, 1989). Estimates suggest that about one-third of nonfatal cases of CO poisoning go undetected and undiagnosed (Abelsohn et al., 2002; Heckerling, 1987; Heckerling et al., 1987). CO poisoning, even when treated with supplemental oxygen (Mathieu and Mathieu-Nolf, 2005), can cause permanent neurocognitive or affective deficits; thus, increased awareness and prevention of CO poisoning is imperative (Weaver, 1999). The mechanism involved in delayed neurological damage after CO exposure was studied in rats. There were significant increases in glutamate release and OH generation during and immediately after CO hypoxia, and CO-exposed rats showed learning and memory deficits that were associated with cell loss in the cortex, globus pallidus, and cerebellum. Both neuronal necrosis and apoptosis were observed, indicating that both necrosis and apoptosis contribute to brain cell death after acute CO poisoning (Piantadosi et al., 1997). This lends some mechanistic support to findings that Parkinsonism, which is an outcome of lesions or losses of dopaminergic neurons, may be associated with exposures to CO (Bleecker, 1988; Choi and Cheon, 1999; Choi, 2002). Necrosis of muscle tissue (myonecrosis) has been reported as possible but fairly unusual sequelae to CO exposure (16–20 cases have been reported in the English-language literature) (Herman et al., 1988; Shapiro et al., 1989; Waisma et al., 1998; Wolff, 1994). Some of the cases involve firefighters, and it is not clear that CO alone is a causal factor. Cyanide, which is a frequent cocontaminant in fires, has been suggested as a contributor to myonecrosis (Shapiro et al., 1989). Marius-Nunez (1990) reported a case of an individual who suffered an acute myocardial infarction (shown by ECG

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and serum enzyme findings) after an acute CO exposure. This case was of interest because the patient’s medical profile was negative for coronary heart disease risk factors, and because a coronary angiogram performed 1 week after admission failed to show coronary obstructive lesions. A similar case was reported (Ebisuno et al., 1986), and the circumstances of both cases suggest that contributing factors to the CO-induced reduction in oxygen supply to the myocardium might include induction of coronary artery spasm, inadequate myocardial perfusion, and a direct toxic effect on myocardial mitochondria. Leikin and Vogel (1986) reported that patients admitted to intensive care units with proven myocardial infarctions had higher COHb levels than a control group, but these differences could have been accounted for by smoking alone, and a relationship to ambient urban CO could not be established. Sokal and Kralkowska (1985) examined 39 cases of acute CO poisoning in Poland. The subjects were poisoned at home by emissions from household gas or coal stoves. The authors found that the duration of CO exposure and the degree of metabolic acidosis, indicated by lactate concentrations in the blood, were better predictors of the clinical severity of symptoms than was the COHb concentration in blood at the time of admission to the hospital. The importance of exposure duration has been suggested in earlier evaluations of CO toxicity and is consistent with the possible involvement of myoglobin in CO toxicology. The prognosis for patients who survive acute CO poisoning is uncertain, particularly in those who develop delayed sequelae after their initial recovery (Deschamps et al., 2003; Ersanli et al., 2004; Gupta et al., 2005; Hwang and Park, 1996; Kanazawa and Yoshikawa, 2004; Kelafant, 1996; Lam et al., 2004; Scheinkestel et al., 1999; Shahbaz Hassan et al., 2003; Webber, 2003). For example, Lee and Marsden (1994) followed 31 patients with CO poisoning sequelae for a year (Lee and Marsden, 1994). Eight had a progressive course and four of the eight died. Twenty-three had a delayed relapse after an initial recovery period of approximately 20 days. Nine of these developed a Parkinsonian state with behavioral and cognitive impairment, but 14 of the cases progressed further and were bed-bound; the deterioration to either condition occurred rapidly over a few days to a week and three died. The mean initial CO hemoglobin level was not different in the two groups. Brain computed tomography (CT) scans were obtained at the onset of sequelae in both groups. Ten patients had a normal CT scan, 13 had white matter low-density lesions, and 4 had globus pallidus low-density lesions. The mechanisms for these sequelae may involve ischemia/reperfusion injury (Mathieu et al., 1996; Wattel et al., 1996) or cerebral biochemical and metabolic changes (Pall, 2001; Thom et al., 2004). 14.6.2

Chronic Exposures

14.6.2.1 Cardiovascular Effects Kristensen (1989) examined the relationship between cardiovascular diseases and exposures in the work environment and concluded that CO exposure increases the acute risk of cardiovascular disease, but that there was no lasting atherosclerotic effect. Stern et al. (1988) performed a retrospective study of heart disease mortality in 5529 bridge and tunnel officers. The socioeconomic and smoking characteristics of the two groups were well matched and the populations were limited to individuals who were assigned their positions and did not transfer between groups. The bridge officers experienced significantly lower CO exposures than the tunnel officers. Significantly elevated risk of coronary artery disease was found in the tunnel officers relative to the bridge officers (61 deaths observed versus 45 deaths expected); however, the risk declined after cessation of exposure, dissipating substantially after 5 years. Although convincing evidence from animal studies is lacking, CO may elevate plasma cholesterol and does appear to enhance


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atherosclerosis when serum cholesterol is greatly elevated by diet (Penney and Howley, 1991). 14.6.2.2 Effects on Lung Function In addition to cardiovascular effects, individuals exposed to relatively high concentrations of CO in both indoor and outdoor environments may also be at risk of lung function decreases. It should be noted, however, that in addition to CO, these individuals were also exposed to high concentrations of other products of combustion and pyrolysis as well, and it is difficult to separate the effects of CO from those of these other compounds, many of which are known to be respiratory system irritants. Firefighters exhibit losses of lung function associated with acute and chronic smoke and CO exposure. Bronchoalveolar lavage fluid from firefighters after smoke exposure shows evidence of inflammation (Bergstrom et al., 1997), and decrements in function (days in which fires were fought compared to routine work shifts without fires) lasted for up to 18 h in some individuals (Sheppard et al., 1986). However, Slaughter et al. (2004) did not find a significant association between CO exposure and pulmonary function deficits in firefighters after exposures during controlled burns. In a study of matched populations of tunnel and bridge officers, whose primary job was to collect tolls, tunnel officers consistently had greater concentrations of COHb, compared to a population of bridge officers with a similar demographic profile that performed essentially similar work. However, the differences were small. Lung function measures of forced vital capacity (FVC) and forced expiration volume in 1 s (FEV1.0) were slightly reduced in tunnel versus bridge officers (Evans et al., 1988). No changes in FVC or FEV1.0 were observed in loggers who complained of dyspnea and eye, nose, and throat irritation after felling trees and cutting logs using chain saws (Hagberg et al., 1985). Exposures to typical ambient concentrations of CO, both outdoors and indoors, have not been significantly associated with pulmonary diseases or lung function decrements (Lebowitz et al., 1987), although other components of ambient pollution do show some significant associations (ozone, particulate matter), as do the use of gas stoves and tobacco smoking. 14.6.2.3 Effects on Pregnancy Outcomes Alderman et al. (1987) performed a case– control study of the association between low birth weight infants and maternal CO exposures in approximately 1000 cases in Denver. CO exposures were assigned to residential locations using fixed-site outdoor monitor data. After controlling for race and education (a surrogate for smoking behavior), no relationship was detected between the assigned CO exposure during the last 3 months of pregnancy and lower birth weights. The investigators suggested that failure to directly account for unmeasured sources of CO exposure, such as smoking, emissions from gas appliances, and exposures to vehicular exhaust, was a limitation of the study design. They also noted that the use of personal monitors for CO would have permitted a more direct evaluation of the potential relationship (exposure evaluations could be made after cases were identified, the relationship of personal to fixed-site assignments could be established, and then applied to the retrospective fixed-site data, author’s note). More recent studies have borne out the association between CO exposure and low birth weight. A 1 ppm average exposure during the first trimester of pregnancy was associated with a 23 g decrease in infant birth weight (Gouveia et al., 2004). Similar results have been obtained in Los Angeles, CA, and Sydney, Australia (Mannes et al., 2005; Ritz and Yu, 1999; Salam et al., 2005). Fetotoxicity has been demonstrated in laboratory animal studies at elevated (125 ppm) levels (Singh and Scott, 1984). Moderate CO exposure can alter neuron development and

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modify neurochemical signaling in rats (Fechter, 1987). Prenatal exposure can adversely alter responses of dopaminergic neurons (Cagiano et al., 1998) and alter development of serotenergic and adrenergic neurons (Fechter et al., 1986; Storm and Fechter, 1985) that can lead to behavioral changes later in life (Fechter and Annau, 1980). 14.6.2.4 Exposure and Relationship to COHb Concentrations A study of over 1500 nonsmoking people sampled as part of the second National Health and Nutrition Examination Survey (NHANES II) demonstrated that CO concentrations measured at fixed-site monitors accounted for only 3% of the variance in blood COHb concentrations, using Spearman rank order correlations between the fixed-site monitor readings and sampled blood COHb concentrations (Wallace and Ziegenfus, 1985). They found that the correlations were not significant (p < 0.05) for 24 of the 36 sampling stations in 20 U.S. cities surveyed. The failure of 8 h average concentrations to correlate strongly with measured COHb concentrations indicates that outdoor monitoring data do not adequately reflect personal CO exposure. Using data from personal monitors worn by a probability sample of over 1500 residents of Denver and Washington, DC, Akland et al. (1985) found that over 10% of Denver residents and 4% of Washington residents were exposed during the wintertime to CO concentrations in excess of 9 ppm for 8 h, or longer. 14.6.3

Controlled Human Studies

Several clinically based studies have been published that have provided a relatively coherent picture of the effects of CO on the cardiopulmonary system. Some of the key studies cited in the 1991 and 1999 CO criteria documents (U.S. EPA, 1991, 1999), as well as those published since then, are described below. 14.6.4

Cardiovascular Effects

Individuals with ischemic heart disease have limited ability to compensate for increased myocardial O2 demands during exercise; hence, exercise testing is often used as a means for evaluating the severity of an individual’s cardiovascular impairment. Four useful parameters of ischemia that are measurable during exercise testing are ST segment depression (at least 1 mVof horizontal or downsloping depression of the ST segment of an electrocardiographic tracing persisting for 70 ms in three successive complexes); exercise-induced angina (chest pain during exercise, which is increased with effort and then resolves with rest—some individuals may experience pain in the jaw, neck, or shoulder areas); impaired work capacity (maximum work levels expressed as a percentage of nomographically predicted, normal values (Bruce, 1971, 1974, 1994); and an inadequate blood pressure response to exercise (blood pressure that falls on exercise (test would be discontinued) or fails to rise more than 15 mmHg at a work level of at least 40% of the predicted norm). There are some individuals who exhibit one or more of these responses during exercise who do not have abnormal coronary arteries, as determined by measuring luminal narrowing using angiographic methods; however, these parameters, taken in combination, can identify 85–90% of people with coronary artery disease (Allison et al., 1996). However, exercise testing alone has limitations with respect to its ability to predict future cardiac events (Fubini et al., 1992). Since CO exposure impairs myocardial O2 delivery, CO exposure would be expected to worsen symptoms of ischemia in individuals with coronary artery disease. Therefore,


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exercise tests of such individuals have been an important means of providing quantitative and dose-related estimates of the potential impact of CO on health. Sheps et al. (1987) exposed 30 subjects with ischemic heart disease, aged 38–75 years, to CO (100 ppm) or air, during a 3-day, randomized, double-blind protocol to achieve an average postexposure COHb concentration of 3.8% on the CO exposure day (COHb on the air exposure day averaged 1.5%). After exposure to either CO or air, subjects performed an exercise stress test. All exercise tests were performed with the subjects in a supine position using a cycle ergometer at the same time of day, with the subjects in a fasting state. The workload was set at 0 for the first min, increased to 200 kpm for the next 4 min, and then increased in 50–100 kpm increments at 4 min intervals until a maximal level was achieved. Exercise was continued until anginal pain required cessation of exercise, fatigue precluded further exercise, or blood pressure plateaued or decreased, despite the increase in workload. All of the subjects were nonsmokers and had documented evidence of ischemic heart disease, defined by exercise-induced ST segment depression (1 mV or more), exerciseinduced angina, or abnormal left ventricular ejection fraction response to exercise (failure to increase 5 units from rest). Not all of the subjects in the study, which included both men and women, reported exercise-induced angina and the CO exposure produced only small, and not significant, decreases in time to onset of angina (1.9%) and maximal exercise time (1.3%) compared to air exposures (Sheps et al., 1987). Times to significant ST decreases, double product (DP; heart rate systolic blood pressure) at significant ST depression, and maximal DP were similar for both air and CO exposure conditions. Double product in the absence of arterial obstructions can be used to estimate myocardial O2 consumption during dynamic exercise (Sim and Neill, 1974). The change in ejection fraction (rest to maximal) was slightly lower for CO exposures (air ¼ 3.5%, CO ¼ 2%; p ¼ 0.049). The authors concluded that there were no clinically significant effects of low-level CO exposures at COHb concentrations of 3.8%. Adams et al. (1988) subsequently extended the above study to an average postexposure COHb concentration of 5.9%, during exercise, using an identical protocol and 30 subjects (22 men, 8 women; mean age 58 years). Not all of the subjects in this study experienced exercise-induced angina, and only 21 subjects reported angina on both exposure days. The time to onset of angina in these 21 subjects was slightly, but not significantly, decreased after CO exposure (10.3%) compared to air exposure. An actuarial analysis of the data, from all subjects reporting angina, indicated that subjects were likely to experience angina earlier during stress on the CO exposure day (p 0.05). The left ventricular ejection fractions at rest were the same after both air and CO exposures; however, the level of submaximal ejection fraction was significantly higher after air, when compared to the CO exposure (3.3%; p 0.05), and the change in ejection fraction, from rest to submaximal exercise, was significantly lower after CO exposure, compared to air exposure (air ¼ 1.6% and CO ¼ 1.2%; p 0.05). No statistically significant exposure-related differences were seen for either maximal ST segment depression, time to onset of significant ST segment depression, or maximal DP. The authors concluded that exposures to CO resulting in COHb concentrations of about 6% significantly impaired exercise performance in subjects with ischemic heart disease. Kleinman et al. (1989) exposed 24 nonsmoking male subjects with stable angina and positive exercise tests to 100 ppm CO or air to achieve an average COHb concentration of 2.9%, during exercise, on the CO exposure day. Subjects ranged in age from 51 to 66 years, with a mean age of 59 years. All but one of the subjects had additional confirmation of ischemic heart disease, such as previous myocardial infarction, coronary artery

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bypass surgery, positive thallium isotope exercise test, or a positive angiogram or cardiac catheterization. Subjects were exposed to CO or to clean air in a randomized, double-blind protocol. Subjects performed an incremental exercise test on a cycle ergometer until the point at which they could detect the onset of their typical anginal pain, and then stopped exercising. Workload was set at 50 W initially and was increased in 25 W increments at 3 min intervals. Blood pressure was measured at the end of each 3 min of exercise, ECG tracings were taken at the end of each minute, and respiratory gas exchange was measured at 15 s intervals and averaged for each minute. Data were analyzed statistically using a twofactor analysis of variance and one-tailed tests of significance. The time to onset of angina was decreased after CO exposure (5.9%; p ¼ 0.046) relative to air exposure. The duration of angina was longer after CO exposure compared to air exposure (8.3%), but this change was not statistically significant. Oxygen uptake at the angina point was slightly reduced after CO exposure compared to air exposure (2.2%; p 0.04), but the increase in O2 uptake with increasing workload was similar on both exposure days. A subgroup of 11 subjects who, in addition to angina, exhibited arrhythmias or ST segment depressions during exercise showed a greater reduction in time to angina after CO exposure, compared to air exposure (10.6%; p 0.016), than did the overall group. The time to significant ST segment depression was significantly reduced for the eight subjects with this characteristic after CO exposure, compared to air exposure (19.1%; p 0.044). The number of subjects exhibiting exerciseinduced ST segment depression identified in this study was small; however, those subjects in whom angina preceded detection of ST segment changes would not have been identified in the protocol used because exercise was stopped at the point of onset of angina. A large multicenter CO exposure study was conducted in three different cities (Allred et al., 1989a, 1989b, 1991). Sixty-three men with documented coronary artery disease underwent exposure to air, 117 ppm CO, or 253 ppm CO, on three separate days in a randomized, double-blind protocol, followed by an incremental treadmill exercise test. Average COHb concentrations of 2.2% and 4.3%, during exercise, were achieved on the two CO exposure days (2.0% and 3.9%, respectively, at the end of exercise). All of the subjects were males, aged 41–75 years (mean age of 62 years), with stable exertional angina and a positive exercise stress test with ST segment changes indicative of ischemia. In addition, all of the subjects had objective evidence of coronary artery disease indicated by at least one of the following:(1) angiographic evidence of at least 70% obstruction in one or more coronary arteries; (2) previous myocardial infarction; and (3) a positive thallium stress test. On each of the exposure days, the subject performed a symptom-limited treadmill exercise test, was exposed to one of the three test atmospheres (clean air, 117 ppm CO, or 253 ppm CO), and then performed a second exercise test. The subjects exercised until the subjects (1) were too fatigued to continue; (2) experienced severe dyspnea; (3) experienced grade 3 angina (on a subjective scale where grade 1 indicated the first perception of angina and grade 4 represented the worst angina the subject had ever experienced); (4) exhibited ECG changes (ST depression 3 mV or important arrhythmias); (5) high systolic (240 mmHg) or diastolic (130 mmHg) blood pressure; (6) a 20 mmHg drop in systolic blood pressure; or (7) a request by the subject. The time to onset of angina and the time to significant ST depression were determined for each test, and the percent changes (preexposure versus postexposure) for the two CO exposure days were compared to the same subject’s response to the randomized clean air exposure. The time to onset of angina was significantly reduced by CO exposure, in a dose-dependent manner (4.2% at 2% COHb, p ¼ 0.054; 7.1% at 4% COHb, p ¼ 0.004). Linear regressions of time to angina versus COHb concentrations for


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each subject indicated that time to angina decreased 1.9 0.8% for every 1% increase in COHb (p 0.01). The time to onset of 1 mV ST segment depression was also reduced by CO in a dose-dependent manner (5.1% at 2% COHb, p ¼ 0.02; 12.1% at 4% COHb, p 0.0001) compared to the clean air exposure. There was a decrease of approximately 3.9 0.6% in time to ST depression for every 1% increase in COHb (p 0.0001). There was a significant correlation between the percent change in the time to onset of angina and the time to onset of ST depression 1 mV ( p 0.0001). There is some evidence that acute hypoxia that can result in myocardial ischemia and reversible angina can also lead to arrhythmias (Dahms et al., 1993; Farber et al., 1990; Jacobs and Nabarro, 1970). Hinderliter et al. (1989) exposed 10 subjects, with ischemic heart disease and no ventricular ectopy at baseline, to air, 100 ppm CO, and 200 ppm CO; COHb concentrations averaged 4% and 6% on the two respective CO exposure days. The exposures were randomized and double blinded. Following exposure, each subject performed a symptom-limited supine exercise test; ambulatory electrocardiograms were obtained prior to exposure, during exposure, during exercise, and over a 5 h postexercise period. The ECGs were analyzed for the frequency and severity of arrhythmias. Eight of the 10 subjects demonstrated evidence of ischemia on one or more of the exposure days (angina, 1 mV ST segment depression, or abnormal ejection fraction response). There were no CO-related increases in the frequency of premature ventricular beats and no multiple arrhythmias occurred. The authors concluded that low-level CO exposure (4–6% COHb) was not arrhythmogenic in patients with coronary artery disease and no ventricular ectopy at baseline. However, researchers from the same team (Sheps et al., 1990) reported on a larger study population (41 subjects) with some evidence of ventricular ectopy, exposed to air, 100 ppm CO, and 200 ppm CO in a similar protocol to that described above. The frequency of single ventricular premature depolarizations (VPDs) per hour increased (p 0.03) from 127 28 (mean SD) after the air exposure to 168 38 after exposure to achieve a COHb concentration of 6%. The frequency of multiple VPDs per hour increased approximately threefold during exercise at 6% COHb, compared to air exposure (p 0.02). No significant differences in these parameters occurred after exposures that achieved COHb concentrations of 4%, compared to air exposures. The subjects who exhibited single VPDs with increased frequency after CO exposure were significantly older than the subjects who had no increased arrhythmias. The subjects who exhibited increased frequencies of multiple VPDs were older, exercised for longer durations, and had higher peak workloads during exercise than those who did not have complex arrhythmias. Leaf and Kleinman (1996) and Kleinman et al. (1998) have also reported evidence of effects of CO exposure on cardiac rhythm after relatively low CO exposures (3% COHb) in a small group of volunteers with coronary artery disease that exhibited abnormal rhythms on one or more exercise test. In all of the above clinical studies of CO-related effects, subjects with coronary artery disease were maintained on individualized regimens of medications, some of which might interact with CO-induced responses, increasing the apparent variations in observed responses. Specifically, blockade of beta-adrenergic receptors (Melinyshyn et al., 1988) and alpha-adrenergic receptors (Villeneuve et al., 1986) was shown to modify hemodynamic responses to CO in animal studies. Examination of the potential influence of medications on observed responses to CO could provide additional insights on the possible mechanisms of action of CO in individuals with coronary artery disease.

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14.6.5

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Cardiopulmonary Effects (Lung Function and Exercise Tolerance)

14.6.5.1 Normal Individuals Reduction of O2 delivery could reduce the ability to perform work in healthy individuals. Studies of the cardiopulmonary effects of CO have demonstrated that maximal O2 uptake during exercise (VO2 max ) decreases linearly with increasing COHb concentrations ranging from 2.3% to 35% COHb, in normals (Horvath, 1981; Horvath et al., 1988b; Shephard, 1984). The linear relationship can be expressed as percent decrease in VO2 max ¼ 0:91 [% COHb] þ 2.2. Changes in VO2 max are significant because they represent changes in an individual’s maximal aerobic exercise (or work) capacity (Ekblom et al., 1975). Klausen et al. (1983) exposed 16 male smokers to CO (5.26% COHb) and compared the effects on maximal exercise performance to performance after 8 h without smoking and performance after smoking three cigarettes (4.51% COHb). Both exposures reduced VO2 max by about 7%, but exercise time was decreased more after cigarette smoking than after CO exposure, suggesting that other components of smoke may contribute to the observed effects (Ekblom and Huot, 1972; Klausen et al., 1983). In a controlled laboratory test, 23 subjects (11 male, 12 female) (Horvath et al., 1988a) were exposed to 0, 50, 100, and 150 ppm CO at four different simulated altitudes (55, 1524, 2134, and 3048 m); following each exposure, an incremental exercise test was performed. COHb concentrations ranged from 0.5 0.2% to 5.6 0.4% of saturation after sea level exposures. The study showed a significant effect of increased altitude on decreased work performance and VO2 max . CO exposure tended to slightly decrease these parameters at all altitudes; however, the statistical analyses did not demonstrate a CO altitude interaction, suggesting that these factors acted independently, and perhaps additively, but not synergistically. The female subjects appeared to be more resistant to the hypoxic effects of altitude than the male subjects. The rate of CO uptake (i.e., formation of COHb) decreased with increasing altitude, in part due to the reduced driving pressure of CO at altitude. In this study, significant fractions of CO were moved to extravascular spaces during exercise, probably in temporary combination with myoglobin, when exercise levels exceeded 80% of VO2 max (i.e., COHb concentrations increased 5 min postexercise compared to concentrations measured at the point of maximum workload). While this might suggest a mechanism in which CO might act in part by directly affecting cardiac myoglobin, evidence for direct cardiotoxicity of CO is still lacking. Horvath and Bedi (1989) demonstrated that long-term, low-level (9 ppm for 8 h) exposures at 2134 m result in lower COHb concentrations than the same exposure at 55 m, again suggesting slower CO uptake during altitude exposure. However, endogenous CO production is increased in rats chronically maintained at high altitudes (1000–6000 m) (McGrath, 1989, 1992), suggesting that high-altitude residents have higher initial COHb concentrations and might therefore achieve 2% or greater COHb levels (the COHb level associated with the CO NAAQS) more quickly than sea level residents. It has been reported that unacclimated workers exposed to about 25 ppm CO at an altitude of 2.3 km above sea level exhibited significantly increased symptoms of headache, vertigo, fatigue, weakness, memory impairment, insomnia, and heart palpitations compared to local residents (Song, 1993). The subjects in these human clinical studies of exercise tolerance have been relatively young and all were in good health. There is not sufficient information available to determine if relationships between CO exposure, altitude, and COHb concentrations would be similar for individuals with coronary artery disease, chronic lung diseases, anemia, or in pregnant women. Kleinman et al. (1998) and Leaf and Kleinman (1996) have demonstrated that hypoxia due to high altitude and CO exposure may cause additive effects on exercise tolerance,


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hemodynamic changes, and cardiologic parameters. The subjects in this study were older men with confirmed coronary artery disease. 14.6.5.2 Individuals with Chronic Obstructive Pulmonary Disease Individuals with COPD usually have limited exercise tolerance because they have low ventilatory capacity, which can result in desaturation of arterial blood and hypoxemia (a relative deficiency of O2 in the blood) and hypoxia (a relative deficiency of O2 in some tissue) during exercise. Exercise performance in such individuals can be improved by providing supplemental O2 (Knower et al., 2001; Kramer et al., 1999). Reduced O2-carrying capacity of blood due to formation of COHb could exacerbate this limitation; hence, individuals with COPD could represent a potentially sensitive group. Aronow et al. (1977) exposed 10 men with COPD, aged 53–67 years to 100 ppm CO for 1 h, achieving increases in COHb from baseline concentrations of 1.4% to postexposure concentrations of 4.1%. Mean exercise time was reduced by 33%. Calverley et al. (1981) exposed six smokers (who stopped smoking 12 h prior to testing) and nine nonsmokers to 200 ppm CO for 20–30 min (increasing COHb concentrations to between 8% and 12% COHb above baseline COHb) and measured the distance each subject walked in a 12 min period. All of the subjects had severe bronchitis and emphysema. Significant decreases in walking distance were seen in individuals with 12.3% COHb or greater (levels that are seen in smokers with COPD). Individuals with severe COPD, even without clinically apparent coronary artery disease, may exhibit exercise-related cardiac arrhythmias. The exercise-induced arrhythmias were associated with arrhythmias at rest, but were not related to the severity of pulmonary disease, O2Hb desaturation, or ECG evidence of chronic lung disease (Cheong et al., 1990). The Sheps et al. (1990, 1991) studies of exercise-related arrhythmias in CO-exposed subjects with coronary artery disease suggest that COPD subjects might be important to study, as well, if they have baseline ectopy. Overall, the information available on individuals with COPD is consistent with the hypothesis that they represent a population potentially at risk of CO-related health effects during submaximal exercise, as may occur during normal daily activities. The available data are, however, based on population group sizes that are too small and too diverse with respect to disease characteristics to draw firm conclusions. 14.6.5.3 Neurotoxicological and Behavioral Effects The neurologic effects of relatively high-level acute CO exposures have been well documented (Gilbert and Glaser, 1959; Lacey, 1981; Remick and Miles, 1977). Subtle neurotoxic effects associated with lower level CO exposures may be underreported or not associated with CO exposure because the symptoms, which resemble those of a flu-like viral illness, may be misdiagnosed (Ares et al., 2001; Balzan et al., 1996; Foster et al., 1999; Raub et al., 2000a). Population-based studies on the potential neurotoxicological and behavioral effects of chronic CO exposure at ambient concentrations have not been reported. However, several clinical studies of CO-related sensory effects that evaluated several different parameters, under controlled laboratory conditions, showed little or no effect at COHb levels up to 17%. Hudnell and Benignus (1989) demonstrated, in a double-blind study, that visual function in healthy, young adult males, as defined by measurements of contrast threshold, luminance threshold, and time of cone/rod break, was not affected by COHb concentrations maintained at 17% for over 2 h. von Restorff and Hebisch (1988) reported no changes in time to dark adaptation and sensitivity after adaptation, at COHb concentrations ranging from 9% to 17%. One earlier study had demonstrated CO-induced visual threshold effects, that is, a slowing of dark adaptation (McFarland, 1973). However, the number of subjects tested was

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small and documentation of the study was scant. Recent studies of temporal resolution of the visual system associated with elevated COHb levels have been reported; however, those involved cigarette smoking studies. The normal capacity for increased blood flow velocity in the central retinal artery in darkness was markedly reduced in smokers, which might explain the reduced dark vision after recent smoking reported in some studies. But this could reflect the combined effects of an increased blood viscosity and the vasoconstrictive action of nicotine in addition to the reduced capacity of the blood to transport O2 due to COHb (Havelius and Hansen, 2005). In general, neurotoxicity at COHb levels near 5% has not been convincingly demonstrated in normal healthy adults (Benignus et al., 1987) even though several early studies had suggested possible effects on critical flicker fusion at COHb levels at or below 9% (Seppanen et al., 1977; Vonpost-Lingen, 1964; Weber et al., 1975). Benignus et al. (1987a) exposed 24 healthy nonsmoking males to 0 or 100 ppm CO for 4 h (mean COHb ¼ 8%). They measured the subject’s ability to perform fast and slow tracking tasks (maintaining the position of a moving point of light on a computer screen using a joystick) and monitoring tasks (judging the brightness of two red spots on a computer screen) once per hour during exposure. CO exposure increased tracking errors, but did not interfere in the monitoring task. An earlier study demonstrated significant decrements in both tracking and monitoring tasks at a COHb concentration of 4.6%, but not at 3.5% (Putz et al., 1979). A large number of studies have investigated the effects of CO on other behavioral parameters; however, effects in general are only seen at COHb concentrations above 5%, and there are inconsistencies among the study results. Other studies, published in 1984 and later, showed interactive effects of exercise and CO exposure (47% COHb) on cognitive tasks (Bunnell and Horvath, 1988), but no changes in visually evoked response potentials in young (23 years) and older (69 years) subjects were observed at 5.3% COHb (Harbin et al., 1988). 14.6.5.4 Fetal Developmental and Perinatal Effects There are both theoretical reasons and supporting experimental data that indicate that the fetus may be more susceptible to the effects of CO than the mother. Fetal Hb has greater affinities for CO and O2 than does maternal Hb. The partial pressure of O2 in fetal blood is about 20–30% of that in maternal blood, because of the greater O2 affinity of fetal Hb. In addition, COHb shifts the O2Hb dissociation curve to the left in maternal blood, reducing the transfer of O2 across the placenta from maternal to fetal circulation. As in adults, the nervous and cardiovascular systems of the fetus are the most sensitive to the effects of CO. For humans, information is available for women who smoked during pregnancy or were acutely exposed to CO; however, most of the available reports do not characterize the relevant CO exposure levels and cannot, in general, rule out toxic effects of cocontaminants. Acute CO exposure may play a role in fetal death (Caravati et al., 1988), and environmental exposures, as well as maternal smoking, have been linked to sudden infant death syndrome in some (Hoppenbrouwers et al., 1981; Hutter and Blair, 1996), but not all studies (Variend and Forrest, 1987). Prenatal exposure to CO affects cholinergic and catecholaminergic pathways in the medulla of the guinea pig fetus, particularly in cardiorespiratory centers, regions thought to be compromised in SIDS (Tolcos et al., 2000). Additional animal studies suggest that high-level maternal CO exposures can have other significant neurotoxicological consequences for the fetus including disruption of neuronal proliferation and possible disruption of markers of neurochemical transmission (Fechter, 1987). Neonatal mortality and low birth weights are more prevalent in children born in high-altitude regions (Moore, 1987; Unger et al., 1988; Yip, 1987), suggesting high-altitude hypoxia interuterine growth, and further suggesting

SUMMARY AND CONCLUSIONS

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that low birth weights in children born to women who smoke during pregnancy could possibly be a result of CO-induced hypoxia. Immune system changes have also been noted in rats exposed to CO prenatally; however, the changes may be reversible (Giustino et al., 1993). 14.6.5.5 CO as a Risk Factor in Cardiovascular Disease Development Evidence from population-based studies indicates that workers exposed to CO in combination with other combustion products from automobile exhaust (Stern et al., 1988) and other workers as well (Kristensen, 1989) have increased risk of development of atherosclerotic heart disease. Also, individuals hospitalized for myocardial infarction frequently exhibit higher COHb concentrations than individuals hospitalized for other reasons (Leikin and Vogel, 1986). Central to the development of atheromatous plaques is the deposition and retention of fibrinogen and lipids within the arterial wall. It is known that cigarette smoke increases the permeability of the arterial wall to fibrinogen. Allen and colleagues (Allen et al., 1989; Allen and Browse, 1990) demonstrated in a canine model that both CO and nicotine in cigarette smoke can produce an atherogenic effect, but they act via different mechanisms. CO increases arterial wall permeability and nicotine reduces clearance of deposited fibrinogen. Activation and dysfunction of blood platelets is associated with production of chemokines that elicit the migration of smooth muscle and inflammatory cells into the vascular intima, which is a major factor in the process of atherogenesis (Munro and Cotran, 1988; Nomoto et al., 1988; Weber, 2005) and in cardiac-related sudden deaths due to the role of platelets in the initiation of thrombosis (Harker and Ritchie, 1980; Meade, 1992). Studies have reported biochemical evidence that cigarette smoking induced both platelet and vascular dysfunctions in apparently healthy individuals (Folts et al., 1990; Krupski, 1991). Production of platelet-derived growth factor (PDGF) by endothelial cells is upregulated in response to hypoxia and is a major growth factor for vascular smooth muscle cells and a powerful vasoconstrictor (Humar et al., 2002; Kourembanas et al., 1990). Platelet dysfunction may also be a contributory cause of thrombosis during pregnancy and may increase fetal mortality and morbidity among women who smoke (Davis et al., 1987). Abnormalities in platelet aggregation occur after CO exposure (Mansouri and Perry, 1982) and may be linked to guanylate cyclase activation (Brune and Ullrich, 1987). When 10 healthy nonsmokers were exposed passively to cigarette smoke (in hospital corridors), resulting in a small increase in COHb concentration from 0.9 0.3% to 1.3 0.6%, before and after passive exposure, respectively (Davis et al., 1989), they showed evidence of changes in platelet aggregation and endothelial cell damage. The changes in endothelial cell counts (preexposure to postexposure) were significantly correlated to changes in COHb concentrations from before to after exposure, but plasma nicotine levels were not. The contribution of CO relative to other components of tobacco smoke in causing platelet dysfunction is not established.

14.7 SUMMARY AND CONCLUSIONS The current CO ambient air standards are designed to protect susceptible individuals from exposures that would result in COHb concentrations of 2% and above. Occupational standards are designed to protect workers from concentrations of 5% COHb. Studies of individuals with coronary artery disease and residents of New York, NY, Denver, CO, Washington, DC, and Los Angeles, CA, suggest that susceptible individuals frequently

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exceed 2% COHb in cities that frequently exceed NAAQS. Control of exposures is difficult because the sources of CO are widespread, the distribution of ambient CO is very nonuniform, and emissions from unregulated sources, especially indoors, probably contribute substantially to individual CO doses. The distributions of COHb concentrations in workers are also very nonuniform but may often reach the 5% level. The contribution of CO to the aggravation of symptoms of myocardial ischemia is reasonably well defined for a selected subset of people with existing coronary artery disease. The individuals comprising the populations tested in the various studies on which this conclusion was drawn were carefully selected to have sufficiently pronounced disease such that effects would be measurable, but they were also sufficiently healthy so that they could perform moderate levels of exercise with minimal risk. Thus, more impaired individuals, who might presumably be at equal or greater risk of detrimental CO-induced health effects, and relatively asymptomatic individuals, so-called silent ischemics, have not been well characterized. Incorporation of broader, possibly more representative, subject populations into the clinical studies of Sheps et al. (1990) and Adams et al. (1988) significantly increased the variance in subject responses and increased the difficulty of attributing statistical significance to observed findings. As shown in Figure. 14.1, there is a reasonable dose– response relationship over the range of 2–6% COHb for the decrease in time to onset of angina in data from five independent studies in which subjects with documented coronary artery disease were exposed to CO and then performed symptom-limited exercise tests. Convincing documentation for effects of CO on other potentially susceptible individuals at ambient exposure levels is becoming available. The most extensive body of evidence of CO effects on pregnant women, fetuses, and neonates comes from the literature on smoking and from acute, high-level accidental CO exposures. In most cases, actual CO exposures are poorly, if at all, documented, and the contribution of copollutants to the observed effects cannot be assessed. However, animal studies demonstrating developmental changes and associations between environmental CO and SIDS indicate that risks to pregnant women, fetuses, and neonates may be important.

FIGURE 14.1 Reduction in time to angina (TTA) following CO exposure in subjects with coronary artery disease. Linear regression shows that TTA is reduced in a dose-dose-dependent manner. Values shown are mean SE.

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The importance of occult CO exposures leading to clinically significant symptoms and effects is becoming well appreciated. The large number of such incidents suggests the potential that there may be many undetected incidents that lead to subclinical manifestations but are ignored if they are not serious enough to prevent relatively normal daily activities. The home environment is very poorly characterized with respect to indoor pollutant levels, and given the apparently large potential for CO-related health effects, the home indoor environment should be the focus of significant new study initiatives. It would seem, from this study, that both occupational and ambient standards are placed near the limits at which significant effects are seen, albeit in sensitive individuals, thus affording a narrow, if any, margin of safety. The number of studies suggesting roles for CO in the development of cardiovascular disease and in infant mortality is increasing, but not yet conclusive. Additional studies under well-controlled conditions, with accurate estimates of CO exposure history, should be a priority.

ACKNOWLEDGMENTS This study was funded in part by the California Air Resources Board and the UCI Center for Occupational and Environmental Health.

REFERENCES Abelsohn A, et al. (2002) Identifying and managing adverse environmental health effects. 6. Carbon monoxide poisoning. Can. Med. Assoc. J. 166(13):1685–1690. Abu-al Ragheb SY, Battah AH (1999) Carbon monoxide fatalities in medicolegal autopsies. Med. Sci. Law 39(3):243–246. Adams KF, et al. (1988) Acute elevation of blood carboxyhemoglobin to 6% impairs exercise performance and aggravates symptoms in patients with ischemic heart disease. J. Am. Coll. Cardiol. 12(4):900–909. Akland GG, et al. (1985) Measuring human exposure to carbon monoxide in Washington, DC, and Denver, Colorado, during the Winter of 1982–1983. Environ. Sci. Technol. 19(10):911–918. Alderman BW, Baron AE, Savitz DA (1987) Maternal exposure to neighborhood carbon monoxide and risk of low infant birth weight. Public Health Rep. 102(4):410–414. Allen DR, Browse NL (1990) The effect of cigarette smoke, nicotine and carbon monoxide on arterial wall permeability and arterial wall uptake of 125I-fibrinogen. Adv. Exp. Med. Biol. 273:95–106. Allen DR, Browse NL, Rutt DL (1989) Effects of cigarette smoke, carbon monoxide and nicotine on the uptake of fibrinogen by the canine arterial wall. Atherosclerosis 77(1):83–88. Allison T, et al. (1996) Cardiovascular stress testing: a description of the various types of stress tests and indications for their use. Mayo Clin. Proc. 71(1):43–52. Allred EN, et al. (1989a) Short-term effects of carbon monoxide exposure on the exercise performance of subjects with coronary artery disease. N. Engl. J. Med. 321(21):1426–1432. Allred EN, et al. (1989b) Acute effects of carbon monoxide exposure on individuals with coronary artery disease. Res. Rep. Health Eff. Inst. (25):1–79. Allred EN, et al. (1991) Effects of carbon monoxide on myocardial ischemia. Environ. Health Perspect. 91:89–132. Amitai Y, et al. (1998) Neuropsychological impairment from acute low-level exposure to carbon monoxide. Arch. Neurol. 55(6):845–848.

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Shahbaz Hassan M, Ray J, Wilson F (2003) Carbon monoxide poisoning and sensorineural hearing loss. J. Laryngol. Otol. 117(2):134–137. Shapiro AB, et al. (1989) Carbon monoxide and myonecrosis: a prospective study. Vet. Hum. Toxicol. 31(2):136–137. Shephard RJ (1984) Athletic performance and urban air pollution. Can. Med. Assoc. J. 131 (2):105–109. Sheppard D, et al. (1986) Acute effects of routine firefighting on lung function Am. J. Ind. Med. 9 (4):333–340. Sheps DS, et al. (1987) Lack of effect of low levels of carboxyhemoglobin on cardiovascular function in patients with ischemic heart disease. Arch. Environ. Health 42(2):108–116. Sheps DS, et al. (1990) Production of arrhythmias by elevated carboxyhemoglobin in patients with coronary artery disease. Ann. Intern. Med. 113(5):343–351. Sheps DS, et al. (1991) Effects of 4 percent and 6 percent carboxyhemoglobin on arrhythmia production in patients with coronary artery disease. Res. Rep. Health Eff. Inst. (41):1–46; discussion 47–58. Sim DN, Neill WA (1974) Investigation of physiological basis for increased exercise threshold for angina pectoris after physical conditioning. J. Clin. Invest. 54(3):763–770. Singh J, Scott LH (1984) Threshold for carbon monoxide induced fetotoxicity. Teratology 30 (2):253–257. Slaughter JC, Koenig JQ, Reinhardt TE (2004) Association between lung function and exposure to smoke among firefighters at prescribed burns. J. Occup. Environ. Hyg. 1(1):45–49. Sokal JA, Kralkowska E (1985) The relationship between exposure duration, carboxyhemoglobin, blood glucose, pyruvate and lactate and the severity of intoxication in 39 cases of acute carbon monoxide poisoning in man. Arch. Toxicol. 57(3):196–199. Solanki DL, et al. (1988) Hemolysis in sickle cell disease as measured by endogenous carbon monoxide production. A preliminary report. Am. J. Clin. Pathol. 89(2):221–225. Song CP (1993) Health effects on workers exposed to low concentration carbon monoxide at high altitude. Zhonghua Yu Fang Yi Xue Za Zhi 27(2):81–84. Stern FB, et al. (1988) Heart disease mortality among bridge and tunnel officers exposed to carbon monoxide. Am. J. Epidemiol. 128(6):1276–1288. Storm JE, Fechter LD (1985) Prenatal carbon monoxide exposure differentially affects postnatal weight and monoamine concentration of rat brain regions. Toxicol. Appl. Pharmacol. 81(1):139– 146. Sue DY, et al. (1988) Metabolic acidosis during exercise in patients with chronic obstructive pulmonary disease. Use of the V-slope method for anaerobic threshold determination. Chest 94(5):931–938. Tan ES, van Veldhuisen DJ, Lie KI (1993) Carbon monoxide poisoning as a trigger for myocardial infarction. Ned. Tijdschr. Geneeskd. 137(44):2266–2268. Thom SR, Keim LW (1989) Carbon monoxide poisoning: a review epidemiology, pathophysiology, clinical findings, and treatment options including hyperbaric oxygen therapy. J. Toxicol. Clin. Toxicol. 27(3):141–156. Thom SR, et al. (2004) Delayed neuropathology after carbon monoxide poisoning is immune mediated. Proc. Natl. Acad. Sci. USA 101(37):13660–13665. Tikuisis P, et al. (1987) A critical analysis of the use of the CFK equation in predicting COHb formation. Am. Ind. Hyg. Assoc. J. 48(3):208–213. Tolcos M, et al. (2000) Chronic prenatal exposure to carbon monoxide results in a reduction in tyrosine hydroxylase immunoreactivity and an increase in choline acetyltransferase immunoreactivity in the fetal medulla: implications for sudden infant death syndrome. J. Neuropathol. Exp. Neurol. 59 (3):218–228.

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15 CHROMIUM Mitchell D. Cohen

15.1 INTRODUCTION Chromium (Cr) is abundant in the Earth’s crust, with both the hexavalent (Cr(VI)) and more predominant trivalent (Cr(III)) forms readily found in nature. Chromite (FeCr2O4) is the most important Cr-containing ore, and is used for production of ferrochromium by direct reduction (Carson et al., 1986). Chemical treatment of chromite, followed by electrolysis, yields Cr metal. Commercially, Cr compounds are commonly used directly in leather/pelt tanning and for electroplating, and as additives in production of pigments, catalysts, corrosion inhibitors, and wood preservatives. Chromium metal is widely used in the steel industry, as a superalloy for jet engines, and for the formation of other alloys. Human exposure to Cr is primarily within the industrial setting or from contact with industrial effluents released into the general environment. Symptoms of acute toxicity include allergic contact dermatitis, skin ulcers, nasal membrane inflammation, and nasal ulceration, while chronic occupational exposure can result in nasal septum perforations, rhinitis, liver damage, pulmonary congestion, edema, and nephritis (Goyer, 1986). Increased incidences of lung and gastric cancers also occur among chronically exposed individuals, while elevations in other types of cancers are also evident (Costa, 1997). The toxicity and carcinogenicity of Cr are largely related to exposure to the metal in its hexavalent state.

15.2 ESSENTIALITY A possible essential role of Cr(III) was demonstrated in 1955 when weanling rats fed a torula yeast-based diet developed small progressive impairments in their glucose tolerance (Mertz and Schwarz, 1955). Subsequent studies with other experimental animals showed that small Cr deficiencies impaired their glucose tolerance, and that the rate of glucose removal was reduced to half its normal value (Schwarz and Mertz, 1959). In addition, severe Cr


529

530

CHROMIUM

deficiencies caused reductions in rodent growth, longevity, fertility, and sperm counts. Correspondingly, there was an increase in glycosuria, aortic plaques, and a rise in the fasting blood glucose and cholesterol levels. Evidence of the beneficial effects of Cr in human nutrition was obtained as a consequence of studies of patients who could no longer ingest food via the normal esophogeal–stomach– intestinal route due to disease or injury. In these patients, surgical implantation of a tube allows for parenteral delivery of fluids containing all essential nutrients. The beneficial effect of Cr was detected as a result of its omission from specially prepared total parental nutrition (TPN) regimens. Two case histories showed that the patients receiving chronic TPN developed considerable weight losses and hypoglycemia (Jeejeebhoy et al., 1977; Freund et al., 1979). These symptoms were reversed with Cr supplementation (as Baker’s yeast) of the parental fluids.

15.3 ENVIRONMENTAL EXPOSURES Occupational exposures to Cr occur during the various stages of its production. Because Cr can be used for many different purposes, there is the potential for exposure in a variety of industries. The most likely risk of occupational exposure is through inhalation of Cr-bearing aerosols. These mixtures are thought to have a wide spectrum of biological activities and are frequently contaminated by other metals (Stern et al., 1984), as well as other known carcinogens such as benzo(a)pyrene. Additionally, there are wide variations in the possible aerosol characteristics, such as the relative proportions of the major oxidation states of the Cr particles, as well as varying solubilities within these fractions (Hertel 1986). Chromium concentrations in soil can range from 0.1–250 ppm Cr and in certain areas, soil content may be as high as 400 ppm Cr (Langard and Norseth, 1979); overall, most soils have been shown to contain on average 50 ppm Cr (Hertel, 1986). However, industrial sources can contribute to significant elevations in the concentration of Cr found in soil. In cases of extensive Cr contamination, such as occurred in 42 Cr-contaminated sites in Hudson County, NJ, concentrations of Cr(VI) and Cr(III) up to 100 and 19,000 ppm, respectively, have been documented in the surrounding soils (ESE, 1989; Paustenbach et al., 1992; Sheehan et al., 1991). Other contaminated industrial sites that were deemed hazardous have included two sites in Odessa, Texas (total soil Cr levels ranging from 720–5000 ppm), and one site each in Woburn, MA (total soil Cr of 1000 ppm), Dixiana, SC (630 ppm Cr), and Vancouver, WA (550 ppm Cr) (U.S. EPA 1984, 1998a and b). Conversely, several agricultural regions throughout the world have been identified as being located upon Cr-deficient soils. This was demonstrated by the fact that both crop yield and quality were improved when Cr was added to the soil. However, it is not clear whether or not the beneficial aspects were due to an effect of the Cr upon the plants themselves, or were a result of interactions of the introduced Cr with other elements or biological agents already present in the soil. While the presence of Cr in phosphate fertilizers is an important source of Cr for crop growth, the downside to the introduction of Cr into the normally Cr-deficient soils also provides a major means for introducing Cr into the environment as a pollutant. Chromium in ambient air originates primarily from industrial sources (i.e., steel manufacturing and cement production) and the combustion of fossil fuels; the content in coal and crude oil varies from 1–100 mg Cr/L and 0.005–0.7 mg Cr/L, respectively (Pacyna, 1986). Airborne particulate matter from coal-fired power plants have been shown to contain Cr in the range of 2.3–31 ppm Cr; however, these levels are reduced to 0.19–6.6 ppm Cr by fly

ENVIRONMENTAL EXPOSURES

531

ash collection processes (Goyer, 1986). Typical atmospheric concentrations of Cr are, on an average, 0.2 to 1.0 ng Cr/m3 in remote continental regions, 1 to 10 ng Cr/m3 for rural and semirural areas, and 13 to 30 ng Cr/m3 in urban areas; other studies have put the latter range at a more expansive 10–100 ng Cr/m3 (see Nriagu and Nieboer, 1988; ATSDR, 2000; Federal Register, 2004), depending on the degree of industrialization. Overall, the distribution of Cr(III) to Cr(VI) is 2:1 in atmospheric Cr emissions; this arises as a result of the fact that most Cr(VI) that enters the air is reduced by the action of many common environmental constituents and other ambient pollutants, including aerosolized acids and dissolved sulfides (Sheehan et al., 1991; ATSDR, 2000). In general, removal of Cr from the atmosphere is the result of either precipitation events or dry deposition. In rural and urban areas, fallout rates for Cr average about 0.2–1.5 and 20– 60 mg Cr/m2/year, respectively (Nriagu et al., 1988); dry deposition rates in areas far away from the point source of emission average between 0.001–0.03 mg Cr/m2/year. The concentrations of Cr in water (U.S. EPA, 1984, 1998a and b) are variable and depend upon salinity. Average concentrations of Cr in American rivers and lakes range from 1–30 mg Cr/L; these values are considerably higher than those found in seawater (0.1–5 mg Cr/L). Drinking water has also been shown to contain higher Cr concentrations than that encountered in river water. For example, in a survey of 84 midwestern cities, the levels of Cr in tap water were found to range from 5–17 mg Cr/L. A controllable source of Cr waste in water is from chrome plating and metal finishing industries, as well as from textile and tanning plants. Industrial wastewater contains total Cr in the range of 0.005–525 mg Cr/L, with concentrations of Cr(VI) averaging from 0.004–335 mg Cr/L. As noted above, most of the Cr that can be encountered in both freshwater and seawater environs is the result of direct deposition of airborne Cr. Oddly, Cr, which is associated with soil, has been deemed not to pose a significant runoff hazard, nor does Cr present much of a threat to aquifers or groundwater supplies as it does not readily leach from soils (U.S. EPA, 1998a and b). This proposed low-risk scenario has been supported by the studies of groundwater Cr levels in the well-studied Hudson County sites; only 1% of the total Cr found in the polluted soils was found to be leachable under stringent extraction procedures (ESE, 1989). The most likely explanation for this was that the majority of the Cr in the soils occurred as water-insoluble Cr(III) (Rai et al., 1986, 1988; Sheehan et al., 1991). However, there may not be a complete absence of a threat from Cr pollution of utilizable water supplies as a result of soil contamination. It might be concluded that under conditions wherein levels of natural reductants might be low in the soil, or the rate of deposition of Cr(VI) onto the soil exceeds that of normal reduction processes, increased amounts of soluble Cr(VI) may penetrate further into the soils and, possibly, reach water-bearing strata. Conditions have been documented in which residents in the vicinity of a Cr-contaminated site were potentially exposed to Cr(VI) in their drinking water at levels up to 10 ppm. For example, in Hinkley, CA, an electric power company pumping station utilizing water coolant laced with potassium chromate routinely discharged the solutions into unlined ponds in the desert. Over time, the Cr permeated into the local aquifer, as well as into wells used for drinking water (Costa, 1997). 15.3.1

Exposure Scenarios

Significant non-occupational Cr exposure of animals and humans is provided through the intake of Cr-containing foodstuffs. The largest sources of dietary Cr are found in meat,

532

CHROMIUM

vegetables, and unrefined sugar, while fruit, fish, and vegetable oils contain fairly small quantities of the metal. Most food, with the exception of herbs and condiments, probably contain less than 100 ppb Cr; concentrations in meat range from 10–60 ppb Cr wet weight (Guthrie, 1982; Kiovistoinen, 1982). Higher Cr concentrations have been measured in some beverages, with typical values in spirits, beer, and wine being 175, 300, and 450 mg Cr/ml, respectively (Jenning and Howard, 1980). Estimated daily intake by adults is between 100–200 mg Cr/day, though large inter-individual variations of 25% of an oral dose (Kuykendall et al., 1996). This variability may relate to the reduction of Cr(VI) to the trivalent form by components of saliva and gastric juice. The oxidation state of Cr is also the determining factor for its transportation via the bloodstream. Trivalent Cr is mainly transported via the serum, bound to the iron-binding transferrin and the b-globulin fraction of serum proteins, however, at high concentrations Cr(III) binds to serum albumin or a1- or a-globulin (Gray and Sterling 1950; Harris 1977). In contrast to Cr(III), Cr(VI) can readily cross the erythrocyte membrane and bind to the globulin portion of hemoglobin following oxidation of the heme group (Gray and Sterling, 1950; Saner, 1980; Nieboer and Jusy, 1986). Inside these cells, Cr(VI) is reduced to Cr(III) by glutathione and then becomes trapped intracellularly. Recent human studies where volunteers ingested Cr(VI) in drinking water showed that a substantial amount enters red blood cells indicated that not all Cr(VI) was reduced to Cr(III) in the GI tract (Kuykendall, 1996). Consequently, the degradation products of erythrocytes may explain, in part, the high concentration of Cr found in the spleen and the slow excretion of Cr from the body. The distribution of Cr from the bloodstream depends on its chemical state. Soluble chelated formsofCrarerapidlycleared,whereascolloidalorprotein-boundformsclearmoreslowly.The latter have a greater affinity for reticuloendothelial system components,such as the liver, spleen, and bone marrow (Hopkins, 1965; Langard, 1982). Accumulation of Cr also occurs in the kidney and testes, whereas retention is less in the heart, pancreas, lungs, and brain. The Cr retained in the liver and kidneys accounts for 45%–50% of total body Cr burdens (Saner, 1980). Excretion of Cr also depends on the oxidation state and occurs primarily via urine and, to a lesser degree, through the feces. Approximately 80% of a parental 51Cr dose is excreted in

TOXICOLOGICAL EFFECTS

535

the urine and 2%–20% in the feces. The biological halftime of 51Cr in humans has been estimated to be 50–60 days (Nieboer and Jusy, 1986). As urinary Cr is generally derived from the dialyzable fraction of serum, postglomerular reabsorption of Cr into renal tubuli results in an intrarenal circulation following exposure (Mutti et al., 1984). However, urinary excretion is generally K2CrO4 > CrO3 > Cr(CH3COO)3 > Cr(NO3)3 (Nakamura and Sayato, 1981). In mammalian cells, while most Cr(VI) salts were mutagenic, Cr(III) compounds produced negative responses. Soluble Cr salts such as K2Cr2O7 and ZnCrO4 have been shown to directly induce gene mutations in Chinese hamster V79 cells (Newbold et al., 1979). The loss of function of target genes after Cr exposure, such as those for resistance to 6thioguanine, ouabain, and 8-azaguanine, has also been documented (Rainaldi et al., 1982). Overall, Cr(III) compounds are relatively nonmutagenic (Langerwerf et al., 1985). However, one study indicated that CrCl3 induced weak mutations in the human fibroblast 6-thioguanine resistance locus, but that this effect was only noted with the insoluble (nonhydrated) form (Biedermann and Landolph, 1990). Determining the degree/mechanism of mutagenicity associated with certain types of Cr– DNA damage (such as DPCs or other adducts that form (mono, binary, ternary)) has been advanced by use of shuttle vectors. Results from these types of studies have shown that while coordination between the phosphodiester backbone of DNA and Cr(III)–GSH or Cr(III)–AA

542

CHROMIUM

complexes created significant premutagenic events (reviewed in Zhitkovich, 2005), there were substantive differences among the types of mutagenic change and the degree of mutagenicity produced. For example, formation of His–Cr(III)–DNA and Cys–Cr(III)– DNA adducts most often resulted in transitions (i.e., GC ! AT) and transversions (GC ! TA); in contrast, Cr(III)–GSH crosslinks primarily produced base substitutions (GC ! TA). Overall, the mutation frequencies induced by the ternary DNA adducts were much greater (GSH > His > Cys > Cr(III) alone) than by binary adducts, and even more so than with monoadducts. Overall, shuttle vector studies have been useful to demonstrate the preferential generation of sequence deletions and point mutations (GC ! AT/TA) in DNA by several reactive Cr species. A study on BHK cells exposed to K2Cr2O7 indicated an inducible inhibition of DNA synthesis. Cr(III)-bound DNA contains novel secondary and tertiary structures that alter DNA polymerase processivity or affect the recognition of bases, thereby resulting in misincorporation (Snow and Xu, 1989; Snow, 1994). This increase in polymerase activity and decreased replicative accuracy was attributed to a Cr(III)-dependent stimulation of polymerase–DNA binding. This effect was more pronounced than the Cr-induced inhibition of either RNA or protein synthesis (Levis et al., 1978). In addition, Cr compounds such as CrO3 and CrCl3 also affected the fidelity of DNA replication in vitro and in intact cells (Sirover and Loeb, 1976; Tsapakos and Wetterhahn, 1983) so that the trivalent Cr-induced introduction of an erroneous base into a replicated strand has little chance of being proofread, excised, and replaced. Besides impacting upon DNA replication and repair mechanisms in vitro, DNA damage in the form of DNA intrastrand breaks and crosslinks as well as DPCs and Cr-DNA adducts have been observed in rats and chick embryo tissues following in vivo exposure to Cr(VI) and not to Cr(III) (reviewed in Standeven and Wetterhahn, 1989). Thus, like Cr(III), Cr(VI) ions can also decrease DNA polymerase replication and fidelity, but the mechanisms are due to a different form of enzymatic competitive inhibition. Unlike Cr(III) ions that complex with nucleotides to form altered substrates that inhibit polymerase activity (Beyersmann and Koster, 1987), Cr(VI) ions directly bind to thiol groups in the enzyme or impart oxidative damage to the enzyme. A number of studies have shown Cr to be capable of inducing chromosomal aberrations and enhancing cell transformation. The morphological transformation of BHK21 cells exposed to Cr(VI) was monitored by the loss of anchorage-independent growth (Hansen and Stern, 1985; Lanfranchi et al., 1988). Similar results were obtained in primary hamster embryo cells treated with K2CrO7 (Hansen and Stern, 1985). In addition, Cr(VI) salts increased the transformation of golden Syrian hamster embryo cells following exposure either in vivo or in vitro. Besides being the cause of direct transformation, Cr(VI) compounds (i.e., CaCrO4, K2CrO4, and ZnCrO4) altered the cell susceptibility to virally induced transformations (Casto et al., 1979). A number of assays have shown that chromosomal aberrations can be induced by both valence states of Cr. Significant increases in the aberrations were observed in cultured BALB/c mouse and Chinese hamster V79 cells exposed to a number of Cr(III) or Cr(VI) salts (Tsuda and Kato, 1977; Newbold et al., 1979; Leonard and Deknudt, 1981; Loprieno et al., 1985). However, sister chromatid exchange was produced in cultured lymphocytes with Cr(VI) salts exclusively (Ohno et al., 1982; Stella et al., 1982). An increase in the aberration frequency in cells obtained from occupationally exposed individuals in Cr plating plants paralleled the observations from the in vitro studies (Bigaliev et al., 1977; Stella et al., 1982). While both valence states of Cr are able to interact with DNA, Cr(III) ions are responsible for decreasing the fidelity of DNA replication. In addition, both Cr(III) and Cr(VI) exhibit a

EXPOSURE GUIDELINES AND STANDARDS

543

clastogenic potency; however, Cr(VI) possesses the greater activity and is also a powerful mutagen in many prokaryotic and eukaryotic cell systems. These properties of Cr(VI) supports the claim that hexavalent compounds are likely to be active carcinogens, although it is more likely that the ultimate species responsible for the carcinogenic/mutagenic effects observed in vivo is the intracellularly derived trivalent form.

15.5 EXPOSURE GUIDELINES AND STANDARDS Since Cr deficiency results in impaired glucose and lipid metabolism, the United States Department of Agriculture (USDA) recommended that a dietary intake of 50–200 mg Cr/day would be safe and adequate in adults (Food and Nutrition Board, NRC, 1980). This range was based on the absence of any symptoms associated with Cr deficiency in a population known to consume an average of 60 mg Cr daily. However, metabolic studies estimated daily intake to be 40) addressing the relationship between diesel exhaust exposure and lung cancer have been reported, if one considers both the studies focusing specifically on diesel exhaust and those including occupations that may be presumed to have received substantial exposures to diesel exhaust based on job title (Lipsett and Campleman, 1999; Bhatia et al., 1998; HEI, 1995). There are only two series of published epidemiological studies where an exposure assessment accompanied the analysis of lung cancer risk. These studies include U.S. railroad workers (Garshick et al., 1987, 1988, 2004, 2006; Larkin et al., 2000; Laden et al., 2006b) and trucking company workers (Steenland et al., 1990, 1992). In both these series of studies, an assessment of current exposure was used to validate the exposure assignments used in the epidemiological analysis. Exposure classifications in other studies were derived from job history by interview of subject or family, job history from employment records, general occupational history from interview, general occupational history from death certificate, and record of membership in a trade union. In the absence of an exposure assessment, limitations in interpreting results from such studies arise from uncertainties regarding the linkage between actual job duties and the extent of diesel exhaust exposure. Although concern has been raised regarding confounding by cigarette smoking in estimates of lung cancer risk, a similarly increased risk is observed in occupational studies whether or not smoking is accounted for, and the air pollution studies assessing lung cancer risk have adjusted for smoking habits. The extent that unmeasured cigarette smoking may confound the assessment of lung cancer risk depends on the association between measures of exposure and smoking. The studies and issues regarding adjusting for smoking were reviewed by the Health Effects Institute in 1995 (HEI, 1995) and in two meta-analyses published in the late 1990s (Lipsett and Campleman, 1999; Bhatia et al., 1998). This section presents results from studies with either large numbers of lung cancer cases or those using the strongest indices of diesel exhaust exposure to provide the reader with an overview of the body of literature that supports an association between lung cancer and diesel exhaust exposure. The studies are summarized in tabular form and divided into studies based on general population registries, occupational studies, and air pollution studies. Selected study characteristics and reported relative risks for lung cancer are presented, and for reports listing multiple study groups or analyses, only the results considered most robust are presented. The 95% confidence intervals (CI) for relative risks are given if they were reported. Hospital, Registry, or General Population-Based Studies American Cancer Society Cohort Boffetta et al. (1988) (Table 16.3) examined the relationship between lung cancer and occupational exposure to diesel exhaust using data from a prospective mortality study begun in 1982 by the American Cancer Society. Living volunteer subjects from across the United States were enrolled in this study by completing a questionnaire that included, among many other items, information on smoking, asbestos exposure, occupation, job held for the longest period, and exposures to diesel exhaust. Cancer Society volunteers checked the status of enrollees every 2 years, and death certificates were obtained for decedents. The analysis was limited to men 40–79 years old at enrollment, whose status was recorded at the end of the first 2-year followup (September 1984). Information was obtained for 11,044 decedents (1266 lung cancer

576

2-year prospective cohort of 461,981 U.S. males aged 40–79 followed over 2 years

French hospital case–control study

Case–control study of FL, LA, NJ hospital and general populations

Benhamou et al. (1988)

Hayes et al. (1989)

Design

Occupation by questionnaire


1,444


Exposure Assessment

1,260

1,266

Number of Lung Cancer Cases

Hospital, Registry, or General Population Based Studies

Boffetta et al. (1988)

References

TABLE 16.3

þ

þ

174

þ

10

112

128

157

48

14

15 5

Number of Cases with Exposure

Control for Smoking

1.5, truck drivers 10 years 2.1, heavy equipment operator 10 years

1.35, transport operators 1.42, motor vehicle operators

1.05, 1–15 years of exposure 1.21, 16þ years of exposure 2.67, miner 2.60, heavy equipment operator 1.59, railroad worker 1.24, truck driver

Relative Risk

0.6–7.1

1.1–2.0

1.07–1.89

1.05–1.75

0.93–1.66

0.94–2.69

1.63–4.37 1.12–6.06

0.94–1.56

0.80–1.39

95% Confidence Interval

577


3,498

Population based case–control study, East and West Germany

Population-based case–control study, Stockholm County, Sweden

Bruske-Hohlfeld et al. (1999)

Gustavsson et al. (2000)

Occupational title

Census classification of occupation

28,744

Hansen et al. (1998) Case–control study, Danish cancer registry

70

Occupation by questionnaire by participant or surrogate

3,792

Case–control study, Detroit cancer registry

Swanson et al. (1993)

Occupation by questionnaire and self-reported exposure

2,584

Case–control study, 18 U.S. hospitals


þ

þ

þ

200

412

716

277 1002

972

121

38

78

114 12

210 0.64–1.09 0.87–6.57

0.75–1.12

0.76, >0–9 years of exposure 1.21, 10–29 years of exposure 1.38, 30þ years of exposure

1.43, all drivers/machine operators 1.84, 10–20 years 1.62, 20–30 years 1.32, 30þ years 1.44, truck, bus, taxi driver West Germany

1.31, truck and bus drivers 1.64, taxi drivers 1.39, unspecified drivers

(continued)

0.97–1.97

0.88–1.65

0.51–1.13

1.34–2.52 1.16–2.24 0.95–1.93 1.18–1.76

1.23–1.67

1.22–2.19 1.30–1.51

1.17–1.46

1.4, heavy truck 0.8–2.4 driver 1–9 years 1.6, heavy truck 0.8–3.5 drivers 10–19 years 2.5, heavy 1.4–4.4 truck drivers 20þ years

0.92, all “probably exposed” 0.83, truck drivers 2.39, self-reported exposure 31 years

578 Number of Lung Cancer Cases Exposure Assessment

Retrospective cohort study; cancer incidence 1971–1995 in Finland

33,664

Census classification of occupation, job exposure matrix

Census 6,266 cases Retrospective classification out of 7,400,000 cohort study; of occupation, person-years cancer incidence job exposure follow-up 1971–1989 in matrix Sweden linked to cancer registry

Design

(Continued)

Guo et al. (2004)


References

TABLE 16.3 Control for Smoking

0.98, lowest exposure in men 1.04, middle exposure in men 0.95, highest exposure in men 0.99, any exposure in men 1.22, any exposure in women

2,436 758 220

3,414 32

1,058

1.2, high probability of exposure 1.3, high intensity of exposure

Relative Risk

1,841


0.96–1.03 0.85–1.73

0.83–1.10

0.97–1.12

0.94–1.03

1.26–1.42

1.10–1.21


HEALTH EFFECTS

579

cases) among 461,981 men, including 174 lung cancer cases among 378,622 men with exposure to diesel exhaust. The asbestos- and smoking-adjusted relative risk for lung cancer of 1.18 (95% CI ¼ 0.97–1.44) was increased among all men with self-reported exposure to diesel exhaust. When the data for all exposed men were stratified by length of exposure to diesel exhaust, the adjusted relative risk for lung cancer was 1.05 (95% CI ¼ 0.80–1.39) for men with 1–15 years of exposure, with a suggestion of an increase in risk among men with at least 16 years of exposure (1.21; 95% CI ¼ 0.94–1.56). Analysis by occupation demonstrated elevated smoking-adjusted relative risks for lung cancer among miners (2.67, 95% CI ¼ 1.63–4.37) and heavy equipment operators (2.60, 95% CI ¼ 1.12– 6.06), railroad workers (1.59; 95% CI ¼ 0.94–2.69), and truck drivers (1.24; 95% CI ¼ 0.93–1.66). Analysis restricted to truck drivers demonstrated no significant difference between relative risks for lung cancer among men reporting diesel exhaust exposure (1.22; 95% CI ¼ 0.77–1.95) and those reporting no exposure (1.19; 95% CI ¼ 0.74–1.89). Analysis by duration of diesel exhaust exposure among truck drivers yielded a positive time-response trend with relative risks of 0.87 for 1–15 years of exposure and 1.33 for 16 years or more of exposure. These results suggest a small positive association between lung cancer risk and occupations with high presumed exposure to diesel exhaust and a trend toward increasing lung cancer risk with time in those occupations. French Case–Control Study Benhamou et al. (1988) reported a case–control study of occupational risk factors among the French population. A total of 1260 male cases observed between 1976 and 1980 were matched by age, hospital of admission, and interviewer with 2084 controls, and information on occupation and smoking was obtained by interview. The smoking-adjusted relative risk for lung cancer among 285 cases and 391 controls was found to be significantly increased (p ¼ 0.01) to 1.35 for transport operators (95% CI ¼ 1.05–1.75) and to 1.42 for motor vehicle drivers (95% CI ¼ 1.07–1.89). National Cancer Institute Pooled Case–Control Study Hayes et al. (1989) reported a case–control study of lung cancer in motor-exhaust related occupations that used data from National Cancer Institute hospital- and population-based studies in Florida (1976–1979), Louisiana (1979–1983), and New Jersey (1980–1981). The study included a total of 1444 male cases and 1893 controls for which occupation and smoking information was obtained by interview. Among 122 cases and 113 controls, the smoking-adjusted odds ratio (OR) for lung cancer was elevated for truck drivers for 10 years or longer (1.5, 95% CI ¼ 1.1–2.0) and heavy equipment operators for 10 years or longer (2.1, 95% CI ¼ 0.6–7.1). Although less specific for diesel exhaust, the OR for lung cancer after 10 years or more of employment in all vehicle exhaust-related jobs was increased to 1.5. In general among job categories, the risk was greater among persons employed 10 or more years as compared to workers employed less than 10 year. American Health Foundation Case–Control Study Bofetta and coworkers (1990) also conducted a case–control study using data from 18 U.S. hospitals that included 2584 male cases of confirmed lung cancer and 5099 controls matched for age, date, and hospital of admission 1977–1987. Information on usual occupation and smoking was obtained by interview, and later questions inquired specifically about diesel exhaust exposure. Data were divided into occupations with no, possible, and probable diesel exhaust exposure, and truck

580

DIESEL EXHAUST

drivers were examined as a subgroup within the probably exposed group. The smokingadjustedORratioforlungcanceramongthegroupwithprobableexposure(210cases)was0.92 (95% CI ¼ 0.75–1.12), and that among truck drivers (114 cases) was 0.83 (95% CI ¼ 0.64– 1.09). The OR increased with duration of service in occupations with probable exposure, reaching 1.49 for 31 years or longer, but the trend did not reach significance (p ¼ 0.18). There wasnosuggestionofanincreasingORratiowithlengthofserviceamongtruckdrivers.Boththis study and the previous American Cancer Society Cohort study by Boffetta and coworkers, although large, were likely to have considerable misclassification of diesel exhaust exposure because classification was based on self-report, reducing the ability to detect an effect of exposure on lung cancer risk. Detroit Area Case–Control Study Swanson et al. (1993) reported lung cancer results from an occupational cancer incidence study of men in the Detroit area using a cancer registry for the period of 1984–1987. Their study involved a total of 3792 cases and 1966 colon/rectum cancer controls, in which work and tobacco use histories were obtained by interview of the subjects or close surrogates. Although numerous occupational groups were listed in the report, those likely to have had the greatest exposure to diesel exhaust were the drivers of heavy trucks. This category included 325 white and 71 black male cases, matched with 164 white and 41 black male controls. The relative risk for lung cancer among white males was related to length of occupation as a driver of heavy-duty trucks, ranging from 1.4 (95% CI ¼ 0.8–2.4) for 1–9 years to 2.5 (95% CI ¼ 1.4–4.4) for 20 years or more. These results suggest an exposure-response relationship, but the results from black men (albeit with fewer cases) did not, although overall lung cancer risk was elevated. Professional Drivers in Denmark Hansen et al. (1998) identified 28,744 men born in 1897–1966 in whom a primary lung cancer was diagnosed in 1970–1989 as identified through the Danish Cancer Registry. Past employment was ascertained by record linkage with a nationwide pension fund that included the dates of starting and stopping work at a particular company. Job titles were retrieved from the Danish Central Population registry. Controls were matched based on year of birth and sex, and had to be alive and employed without cancer before the case was diagnosed. Adjusting for socioeconomic status based on occupational title, the OR for truck and bus drivers (972 cases and 668 controls) was 1.31 (95% CI ¼ 1.17–1.46), for taxi drivers (277 cases and 149 controls) OR ¼ 1.64 (95% CI ¼ 1.22–2.19), and for unspecified drivers (1002 cases and 598 controls) OR ¼ 1.39 (95% CI ¼ 1.30–1.51). For both truck and bus drivers and taxi drivers, there was an increasing risk of lung cancer with greater years of work. Data on smoking in these occupations was available indirectly from national surveys conducted in Denmark in 1972 and 1983, and smoking rates were similar among drivers and nondrivers in working men, suggesting that the results were not confounded by smoking. Occupational Exposure to Diesel Engine Emissions in Germany Bruske-Hohlfeld et al. (1999) studied the association between occupation and lung cancer in diesel exposed workers in a pooled analysis of two case–control studies that included 3498 male cases and 3541 population controls. Information regarding occupational exposure and smoking was obtained by questionnaire, and jobs were divided into four groups: professional drivers of trucks, buses, and taxies; other exposed jobs, including diesel locomotive and diesel forklift truck drivers; machine operators, including bulldozer, grader, and excavator drivers; and tractor drivers. There were 1146 men occupationally exposed to diesel exhaust (716 cases and

HEALTH EFFECTS

581

430 controls). For all jobs, the crude risk of lung cancer was 1.91, which was reduced to 1.43 (95% CI ¼ 1.23–1.67) by adjusting for smoking and asbestos exposure. There was an increased lung cancer risk with greater years of exposure through 10–20 years (OR ¼ 1.84; 95% CI ¼ 1.34–2.52), but risk slightly decreased with 20–30 years (OR ¼ 1.62; 95% CI ¼ 1.16–2.24) and greater than 30 years of exposure (OR ¼ 1.35; 95% CI ¼ 0.95–1.93). Among those with other diesel exposed jobs (99 cases, data not shown) lung cancer risk was also elevated (OR ¼ 1.53; 95% CI ¼ 1.04–2.24), with evidence of greater risk with greater years of exposure; for tractor drivers (52 cases) there was evidence of increasing risk with increasing duration of employment, and there was an overall increased risk in heavy equipment operators (81 cases; OR ¼ 2.32; 95% CI ¼ 1.44–3.70). Differences in the risk for truck, bus, and taxi drivers were examined separately for East Germany and West Germany. In West Germany (412 cases), the overall risk for professional drivers of trucks, buses, and taxis (data not shown) was 1.44 (95% CI ¼ 1.18–1.76), whereas in East Germany (122 cases) it was 0.83 (95% CI ¼ 0.60–1.14). In West Germany, there was a suggestion of an increase with greater driving hours, and a greater risk if one started to drive 1946, and in particular, 1956 (OR ¼ 1.60; 95% CI ¼ 1.32–1.96), which most likely represented times with the greatest numbers of vehicles on the rounds, particularly diesel cars and trucks. The differences in between East Germany and West Germany were attributable to differences in traffic density, which was estimated to be five times greater in West Germany. Occupational Factors and Lung Cancer Risk in Sweden Gustavsson et al. (2000) studied lung cancer identified in cancer registries in men aged 40–75 who were residents of Stockholm County, Sweden, between 1985 and 1990, and who had lived outside the county for no more than 5 years between 1950 and 1990. Referents were chosen at random from the general population, and included mortality-matched referents. Smoking and occupational histories were obtained from the subject if alive, or from next of kin. Occupations were coded into an exposure intensity and probability matrix, and job-specific historical values of exposure to NO2 were used to estimate occupational exposure to diesel exhaust. Residential exposures to NO2 were modeled using a historical inventory linked to a traffic grid. Adjusting for smoking and ambient NO2, there was increased lung cancer in the highest cumulative occupational diesel exhaust exposure category (1.63; 95% CI ¼ 1.14–2.33), and for 30 years of exposure relative risk ¼ 1.38 (95% CI ¼ 0.97–1.97) (data not shown). Swedish Cancer Environment Registry Study Boffetta et al. (2001) investigated the risk of cancer among workers exposed to diesel exhaust using the Swedish Environment Register III that contains nationwide data on cancer incidence for 1971–1989 and linked this to occupation and industry of employment as reported in the 1960 census. Using a jobexposure matrix, exposures were graded based on intensity and probability of exposure to diesel exhaust. Mortality was ascertained between 1971 and 1989 by linkage to the Swedish Cancer Registry and Register of Causes of Death. There were a total of 28 million person-years of observation, for which 26% was in people classified as exposed. The analysis of lung cancer risk in men (6266 cases overall) with a high probability of exposure was increased (1.2; 95% CI ¼ 1.10–1.21), and there was an increasing risk with increasing exposure intensity (high intensity relative risk ¼ 1.3; 95% CI ¼ 1.26– 1.42). There were fewer cases (n ¼ 57) and less exposure in women, and lung cancer risk was not increased. Although there was no specific information regarding smoking, it was noted that other smoking-related tumors were not increased based on diesel exposure, suggesting no confounding by smoking.

582

DIESEL EXHAUST

Finnish Worker Study Guo and coworkers (2004) studied the association between lung cancer and occupation during 1971–1995 in Finland by linking census occupation reported in 1970 to lung cancer mortality. A job-exposure matrix was used to estimate exposures to gasoline and diesel exhaust based on occupation. The authors used estimates of NO2 exposure as surrogates of diesel exposure and estimates of CO exposures as measures of gasoline engine exhaust exposures. There were 33,664 cases of lung cancer identified. Based on exposure to NO2, the overall relative risk for lung cancer was 0.99 (95% CI ¼ 0.96–1.03) among men and 1.22 (95% CI ¼ 0.85–1.73) among women. Based on estimated CO exposures (data not shown), the relative risk of lung cancer was 1.05 (95% CI ¼ 1.01– 1.09) among men and 1.61 (95% CI ¼ 1.23–2.12) among women. Smoking information was obtained from general population surveys and was incorporated into the analysis, but the specific methods were not stated. Results based on job title were not reported. The authors attributed the generally negative results to low exposures due to the operation of diesel vehicles and other equipment in rural areas with a low population density. It is also possible that considerable misclassification was introduced by the quantitative estimation of exposures. Occupational Case–Control And Cohort Studies Swedish Dockworker Studies Gustafsson et al. (1986) (Table 16.4) compared the incidence of lung cancer among male Swedish dock workers to that among the Swedish male population. Diesel trucks were introduced into Swedish ports in the late 1950s and became prevalent during the 1960s. The cohort consisted of 6071 men employed for a minimum of 6 months before 1974 and followed from 1961 to 1981. Twenty percent of the cohort had 30 or more years of service, and only 10% had less than 5 years of service. There were 70 cases of lung cancer among the 1062 cohort deaths. The relative risk for lung cancer among the dock workers was found to be significantly increased to 1.32 (95% CI ¼ 1.05–1.66). Similarly, Emmelin et al. (1993) reported increasing relative risks for lung cancer with exposure time for 50 cases among nonsmoking Swedish dock workers ranging from 1.0 (no increase) for the lowest category to 2.9 for the highest. Exposure to diesel exhaust was estimated based on diesel fuel consumption and number of workers in each Swedish port, and the cases and controls were selected from male dock workers employed for at least 6 months during 1950–1974, with case ascertainment starting in 1960 through 1982. Based on 50 cases and 154 referents with complete information available and adjusting for smoking (yes/no), there was an increase in the OR ratio for lung cancer with increasing exposure for three indices of exposure (years since diesel equipment was used in a port, estimates of cumulative fuel consumption, and years that fuel use was above a minimum level in a port), consistent with an exposure-response relationship. U.S. Teamster Case–Control Study Steenland et al. (1990, 1992) reported a study of lung cancer among truck drivers in the Central States Teamsters’ Union. The study population included a total of 996 cases and 1085 controls, for whom death certificates were obtained, and information on work history, smoking, and asbestos exposure was obtained from next of kin. Job history information was also available from the worker’s retirement applications. The subjects died in 1982–1983 and were receiving pensions, which required a minimum of 20 years of union membership. Covariates included in the analysis were age, smoking, asbestos exposure, and jobs with diesel exposure. Heavy-duty diesel

583

Retospective cohort of 6,071 male dockworkers 1961–1981

Nested case–control study in dockworker cohort

Case–control study of mortality based in Teamster Pension fund

Emmelin et al. (1993)

Steenland et al. (1990)

Design

Use of diesel equipment and fuel consumption Occupation by questionnaire next of kin questionnaire, retirement application

996

Job title

Exposure Assessment

50

70


Occupational Case–Control and Cohort Studies

Gustafsson et al. (1986)

References

TABLE 16.4

162

þ

37

36

213

1,002

228

12 19 19

70


þ

Control for Smoking

1.06, 1–11 years intercity truck driver 1.41, 12–17 years intercity truck driver 1.55, 18þ years, intercity truck driver 1.11, 1–11 years city truck driver 1.15, 12–17 years city truck driver 1.79, 18þ years, city truck driver Years of work based on years after 1959, the date diesel trucks were introduced

1.0, reference 1.6, medium exposure 2.9, highest exposure

Standardized mortality ratio ¼ 1.32

Relative Risk

(continued)

0.94–3.42

0.80–4.19

0.94–3.42

0.97–2.47

0.90–2.21

0.68–1.70

0.83–1.10 0.5–5.1 0.8–10.7

1.05–1.66


584 Occupation by questionnaire

122

Retrospective cohort study of 389,000 male Swedish construction workers 1971–1993

Jarvholm and Silverman (2003)

Occupational title

38

Retrospective cohort mortality study of 5,536 German Potash miners 1970–1994

Saverin et al. (1999)

Employment in cohort

Exposure Assessment

Occupational title

Nested case–control study in 1998 cohort

Soll-Johanning et al. (2003)

473


153

Retrospective cohort mortality study of 18,174 bus drivers and other tramway employees in Copenhagen 1990–1994

Design

(Continued)

Soll-Johanning et al. (1998)

References

TABLE 16.4

Indirect assessment

Indirect assessment

þ

Control for Smoking

61

61

38

11

153

473


0.87, heavy equipment operator 1.29, truck driver Risk compared electricians/carpenters

2.17, exposed production workers 1.68, based on highest category of cumulative exposure

0.97, years of work as bus driver

1.2, compared to population rates

Relative Risk

0.99–1.65

0.66–1.11

0.49–5.8

0.79–5.99

0.96–0.99

1.1–1.3


585

Retrospective cohort mortality study in 43,826 Canadian railroad retirees

Case–control study of all active and retired U.S. railroad workers (650,000 workers). Deaths collected 1981–1982

Retrospective cohort mortality study of 54,973 railroad workers 1959–1996

Retrospective cohort mortality study of railroad workers 1959– 1996, includes 39,388 deceased workers

Howe et al. (1983)

Garshick et al. (1987)

Garshick et al. (1988, 2004)

Garshick et al. (2006)

Yearly job title from retirement board. Next of kin smoking history Yearly job title

Yearly job title

4,351

4,055

Job title at retirement

1,256

933

Indirect adjustment; imputation of smoking history using information from 1987 case– control study

Indirect adjustment

þ

1.40, train crews, unadjusted for smoking Estimated smoking adjusted relative risk 1.17–1.27 1.35, train crews, unadjusted for smoking 1.22, train crews, smoking adjusted

2,358

1.41, for 20 years of work in an exposed job

1.20, possible exposure 1.35, probably exposed

2,479

335

374 306

1.24–1.46 1.12–1.32

1.30–1.51

1.06–1.88

p ¼ 0.012 p < 0.001

586

DIESEL EXHAUST

trucks for inter-city use were introduced in the industry during the 1950s. Using retirement application job history, drivers of inter-city trucks (552 cases and 604 controls) with the greatest duration of work after 1959 (18þ years) had the greatest lung cancer risk (OR ¼ 1.55; 95% CI ¼ 0.97–2.47) with a significant linear trend with years of work. However, for city drivers who drove gasoline-powered heavy-duty trucks (113 cases and 135 controls), the risk of lung cancer was similarly elevated with 18þ years of work after 1959 (OR ¼ 1.79; 95% CI ¼ 0.94–3.42). City drivers with shorter job duration had lower risks. These data are consistent with the exposure assessment that accompanied the study (Zaebst et al., 1991; Steenland et al., 1992), indicating that local city truck drivers and inter-city truck drivers had similar EC exposures that reflected background exposure from traffic rather than exposure from their own truck. Danish Urban Bus and Tramway Study Soll-Johanning et al. (1998) conducted a retrospective cohort mortality study of 18,174 bus drivers or tramway employees in Copenhagen during 1990–1994 and reported an increased risk of lung cancer based on linkage to the Danish Cancer Registry. Among workers employed for 3 or more years, the relative risk of lung cancer compared to Copenhagen rates was 1.2 (95% CI ¼ 1.1–1.3) based on 473 cases. The same group (Soll-Johanning et al., 2003) conducted a nested case–control study using these cases. There were 257 lung cancer cases where information regarding smoking history was potentially available from the wife or case file. This information was obtained for 153 cases, and the study also included 351 controls. An increased risk of lung cancer was not related to duration of employment, and there was a decreased risk of lung cancer with increasing years or work as a bus driver. There was some limited information about driving route, and driving in areas of high air pollution was not associated with risk. Lung Cancer Mortality in Potash Miners Saverin et al. (1999) studied lung cancer mortality in Potash miners in Germany. Diesel equipment was introduced into Potash mines in 1969–1970, and in 1991, the mines closed. Workers had medical examination every other year, and records on smoking were maintained through 1982. There were 5536 men who had worked underground for at least 1 year after 1969, and mortality was ascertained for 1970– 1994. Estimates of diesel exposure were obtained in 1992 and expressed as total carbon in respirable dust, and because technology had not changed, these levels were assumed to be representative of previous exposure. Although medical records were found to classify former smokers as nonsmokers in 28% of cases when compared to a personal interview, smoking was not associated with exposure so was not considered to be a confounder in the analysis. The exposed workers were the production workers, and the relative risk of lung cancer in production workers who had worked underground for at least 10 years (11 cases) compared to other workers (6 cases) was elevated but imprecise (relative risk ¼ 2.17; 95% CI ¼ 0.79– 5.99). Results using categories of cumulative exposure gave similar results (relative risk ¼ 1.68; 95% CI ¼ 0.49–5.8). Although miners typically have much higher diesel exhaust exposure levels than other occupations, the cohort was small and exposure for many was of short duration, potentially limiting the detection of lung cancer risk. Swedish Truck Driver and Construction Vehicle Case–Control Study Jarvholm and Silverman (2003) studied lung cancer in 389,000 male Swedish construction workers and identified truck drivers and heavy equipment operators. Workers were identified based on health examinations during 1971–1993 and linked to the Swedish National Cancer Registry and National Death Registry and lung cancer cases identified through 1995. Subgroups of

HEALTH EFFECTS

587

exposure were created based on whether a cabin was on the construction equipment. There were 14,364 heavy equipment operators (61 lung cancer cases), 6364 truck drivers (61 lung cancer cases), and 119,984 carpenter/electrician referents (512 lung cancer cases). Eighty percent of the heavy equipment operators operated the same machine between the first and last examination, indicating high job stability. Heavy equipment operators had a lower incidence of lung cancer compared to electricians/carpenters (standardized incidence ratio; SIR ¼ 0.87; 95% CI ¼ 0.66–1.11) and the general population (SIR ¼ 0.76; 95% CI ¼ 0.58–0.97), whereas truck drivers had a greater incidence of lung cancer compared to electricians/carpenters (SIR ¼ 1.29; 95% CI ¼ 0.99–1.65) and the general population (SIR ¼ 1.14; 95% CI ¼ 0.87–1.46), and mortality ratios were similar. Information on smoking habits was available from the examination in a subset of workers, and smoking rates were similar among the jobs, making it unlikely that differences in smoking rates accounted for the finding. When the heavy equipment operators were categorized based on use of cabins, lung cancer risk for never in a cabin was 0.86 (95% CI ¼ 0.5–1.6); for sometimes in a cabin, SIR ¼ 0.71 (95% CI ¼ 0.5–1.0); and for always in a cabin, SIR ¼ 0.50 (95% CI ¼ 0.20–0.70). This trend (p < 0.001) was suggested that working inside a cabin while on a construction vehicle was associated with a lower lung cancer risk. Railroad Worker Case–Control and Cohort Studies Howe et al. (1983) conducted a retrospective cohort study of lung cancer among 43,826 male employees of the Canadian National Railway retired and alive in 1965 or retiring between 1965 and 1977. The total of 16,812 deaths included 933 deaths from lung cancer. The subjects were classified by job at the time of retirement into nonexposed, possibly exposed, and probably exposed to diesel exhaust. A highly significant relationship was found between relative lung cancer risk and the presumed level of exposure:nonexposed ¼ 1.00, possibly exposed ¼ 1.20, and probably exposed ¼ 1.35. As noted earlier, the U.S. railroad industry converted from steam to diesel powered locomotives mainly starting in the late 1940s, and by 1959, 95% of the locomotives in service were diesel powered. Garshick and coworkers (1987) collected death statistics over 1 year (1981–1982) from a population base of 650,000 active and retired male U.S. railroad workers with 10 years or more of service, using records from the Railroad Retirement Board. Their study included a total of 1256 exposed cases and 2385 controls, assigned on the basis of job records and contemporary measurements (early 1980s) of diesel exhaust concentrations in similar job environments (Woskie et al., 1988a, 1988b). Lung cancer cases and controls were matched by birth and death date. The cases and controls were classified by age, length of service, smoking (next of kin history), and likely exposure to asbestos. Exposed workers included train crews and locomotive shop workers; unexposed workers included workers not in these job groups. The cases were divided at age 64 into younger and older groups, with the younger group presumed to have more years of diesel exposure because of the dates of railroad dieselization, and years of work starting in 1959 in a diesel-exposed job was used as a continuous exposure variable. After adjustment for smoking (pack-years) and asbestos exposure (yes/no), the odds ratio for lung cancer among 335 cases and 637 controls for working 20 years in an exposed job was 1.41 (95% CI ¼ 1.06–1.88) (data not shown). After similar adjustments, the odds ratio for workers with 20 or more years of diesel exposure was 1.64 (95% CI ¼ 1.18–2.29), and in analyses examining mortality in the train crews with 20 or more years of exposure was 1.55 (95% CI ¼ 1.09–2.21). To exclude the effects of recent diesel exhaust exposure, the data were analyzed excluding exposures during the 5 years preceding death, and the relative risk for lung cancer remained similarly elevated.

588

DIESEL EXHAUST

Garshick et al. (1988, 2004) also conducted a retrospective cohort study of lung cancer mortality among 54,973 white male U.S. railroad workers, 40–64 years old in 1959, who had begun work 10–20 years earlier. In the original publication, mortality was assessed through 1980 (Garshick et al., 1988), and then later updated through 1996 (Garshick et al., 2003). The cohort was selected on the basis of job title in 1959 using records from the U.S. Railroad Retirement Board. As in the case–control study, exposed workers included train crews and locomotive shop workers, and the unexposed group included clerks, ticket and station agents, and signalmen. Jobs with the most likely exposure to asbestos were excluded. There were 45,593 deaths over the 38 years of followup, including 4351 lung cancer deaths. Workers on operating trains (train crews) had a relative risk of lung cancer mortality of 1.40 (95% CI ¼ 1.30–1.51). There was no increase in lung cancer mortality with greater years of work that was attributed to a healthy worker survivor effect and insufficient information regarding historical changes in railroad exposures. Locomotive shop workers did not have an increased lung cancer risk, but it was later noted that the job titles of the workers included were not specific for diesel locomotive shops, thereby reducing the ability to detect an effect of exposure. Although there was no specific cigarette smoking history information available from the workers in the cohort, smoking information was available from surveys of railroads workers, including the 1987 case–control study conducted by Garshick et al., 1987; Larkin et al., 2000). This information was used to estimate an effect of smoking that reduced the relative risk to 1.17–1.27 (Garshick et al., 2004). The authors also conducted an additional analysis to assess whether differences in smoking behavior between diesel exposed and unexposed workers influenced the risk of lung cancer (Garshick et al., 2006). A simulation of smoking behavior using cause of death, birth cohort, age, and job-specific smoking prevalence from the 1987 lung cancer case–control study was conducted for 39,388 deceased railroad workers. The risk of lung cancer among exposed workers unadjusted for smoking was 1.35 (95% CI ¼ 1.24–1.46), and after adjustment an excess risk remained (1.22; 95% CI 1.12–1.32). In order to improve the estimation of historical exposures during the transition from steam to diesel locomotives, historical information on diesel locomotives used by each railroad was obtained (Laden et al., 2006b). Starting in 1945, annual railroad-specific weighting factors for the probability of diesel exposure were calculated. Among workers hired after 1945, as diesel locomotives were introduced, the relative risk of lung cancer for any exposure was 1.77 (95% CI ¼ 1.50–2.09), and there was evidence of an exposureresponse relationship with exposure duration. Air Pollution Cohort Studies Diesel exhaust contributes to ambient NO2 exposures, and two studies have used estimated values of NO2 as an index of overall traffic exposure and have assessed its relationship with lung cancer. Nyberg et al. (2000) conducted a case– control study among men 40–75 years old that included all cases of lung cancer 1985–1990 in residents of Stockholm County, and smoking histories were obtained from next of kin for 1042 cases and 2364 controls. Geographic Information System techniques were used to estimate residential exposures to NO2 as an index of traffic exposure by linking a regional emission database to a road network. Adjusting for smoking, age, occupational exposures to diesel exhaust, socioeconomic group, and asbestos, the relative risk of lung cancer for the highest decile of average NO2 exposure lagged 20 years was 1.44 (95% CI ¼ 1.05–1.99). Nafstad et al. (2003) conducted a similar study in Oslo, Norway, linking 16,209 participants

HEALTH EFFECTS

589

in a prospective health study in 1972–73 to estimates of outdoor NO2 levels at their residential address. Adjusting for age, smoking, and education, there was a significant increase in lung cancer risk (1.08; 95% CI ¼ 1.02–1.15) for every 10 mg/m3 increase in NO2. In contrast, a study of cancer incidence and residential traffic density in Amsterdam 1989– 1997 did not find a consistent association between distance from a roadway, traffic density, and lung cancer (Visser et al., 2004). Two air pollution studies have related fine PM (PM2.5) to lung cancer mortality but were not able to examine more specific markers of traffic or diesel exposure. In the American Cancer Society Prevention II follow-up study, 1.2 million adults enrolled in 1982 had mortality ascertained through 1998 (Pope et al., 2002). Using national networks of monitors, average PM2.5 levels were linked using zip code to 319,000 people in 51 metropolitan areas. Adjusting for age, sex, race, smoking, education, and diet, there was an increased risk of dying of lung cancer (1.14; 95% CI ¼ 1.04–1.23) for every 10 mg/m3 increase in PM2.5. Similar results were obtained from the Harvard Six Cities Study where lung cancer mortality in 8096 participants starting in the 1970s through 1990 was related to average PM2.5 levels (relative risk ¼ 1.27; 95% CI ¼ 0.96–1.69), adjusting for smoking and multiple other covariates (Laden et al., 2006a). Summary of Epidemiological Evidence for Lung Cancer Despite limitations in the assessment of exposure, when considered together, the weight of the human epidemiologic evidence reviewed above supports a small increased risk of lung cancer associated with diesel exhaust exposure. This increased risk was observed in workers with long-term employment in a variety jobs involving exposures to diesel exhaust. The studies indicating statistically significant increases gave estimates of increases ranging from approximately 20% (relative risk of 1.2) to approximately two-fold increases (relative risk of 2.0). The level of confidence with which one can draw conclusions from the epidemiological studies of workers regarding cancer risks from lower exposures is limited due to lack of historical exposure estimates. Historical exposures were probably higher than current exposures, and there is also uncertainty regarding whether the available measurements were representative of industry-wide exposures. The existing data do not allow determination of either (1) the magnitude of increase per unit exposure; or (2) a description of the exposure-response relationship, with a high level of confidence. The EPA and others have presumed that the health risk is present at ambient levels because exposures experienced by some occupations overlap with general population exposures, such as in professional drivers. Results from general air pollution studies also support the plausibility of a lung cancer risk from ambient PM2.5 exposures that include DPM as a minor mass component. The dose required, the specific carcinogenic agent in diesel exhaust, and the extent to which the cancer risk is unique to diesel exhaust remain uncertain. Two ongoing occupational epidemiological studies may provide improved knowledge about the exposure-response relationship for lung cancer. A national exposure assessment of the trucking industry exposures was recently performed, and the assessment was specifically designed to be linked to an epidemiologic database examining lung cancer mortality (Smith et al., 2006; Davis et al., 2006). The analyses in this study are still in progress. A study of the relationship between exposure and lung cancer among workers in metal and nonmetal mines conducted by the National Cancer Institute (NCI) and the National Institute for Occupational Safety and Health (NIOSH) is also still in progress (Monforton, 2006). At this time, because lung cancer takes years to develop, and in the absence of a long-term prospective environmental

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epidemiological study accompanied by measurements of exposure to diesel exhaust, estimates of lung cancer risk due to diesel exhaust or other traffic emissions remain dependent on historical exposure estimates. 16.4.1.3 Epidemiological Evidence for Bladder Cancer The bladder is the only organ, other than the lung, for which there has been significant concern for cancer associated with diesel exhaust exposure. There have been several studies of the incidence of bladder cancer among populations presumed to be heavily exposed to diesel exhaust. A few studies focused primarily on the relationship between bladder cancer and diesel exhaust exposure, but most included the bladder in the context of broader cancer surveys. Most of the difficulties described above for the studies of lung cancer are also inherent in the studies of bladder cancer. These include lack of quantification of exposure, lack of uniqueness of the exposure materials to diesel exhaust, the presence of several confounding factors, and potential exposure misclassification. Information on the association between diesel exhaust exposure and bladder cancer was reviewed by Boffetta and Silverman (2001), who identified 35 studies and performed a meta-analysis by job group. Among 15 studies where there was occupational exposure as a truck driver, relative risk ¼ 1.17 (95% CI ¼ 1.06–1.29), among 10 bus driver studies, relative risk ¼ 1.33 (95% CI ¼ 1.22–1.45), among heavy equipment operators in 5 studies, relative risk ¼ 1.37 (95% CI ¼ 1.05–1.81), and using a job-exposure matrix in 10 studies, relative risk ¼ 1.13 (95% CI ¼ 1.00–1.27). In contrast, Boffetta et al. (2001) also investigated the risk of bladder cancer attributable to diesel exhaust exposure in a study where national data on cancer incidence during 1971–1989 in Sweden were linked to occupation and industry as reported in the 1960 census. Using a job-exposure matrix, exposures were graded based on intensity and probability of exposure to diesel exhaust. Mortality was ascertained between 1971 and 1989 by linkage to the Swedish Cancer Registry and Register of Causes of Death. This study included 3669 cases of bladder cancer classified as low, medium, or high intensity exposure and 12,287 cases without exposure. In this very large study, no association with diesel exhaust exposure was observed. A similar study was conducted in Finland (Guo et al., 2004) that included 771 cases with exposure and 4314 cases without exposure and no association was observed. The weight of the present evidence suggests that there may be a small positive risk for bladder cancer among truck drivers and other long-term workers in occupations presumed to be exposed to diesel exhaust, but as illustrated by the latter two studies, the epidemiologic results are not as consistent as for lung cancer. 16.4.1.4 Current Classifications of Cancer Risk Several health and regulatory organizations have reviewed the evidence for human carcinogenesis from inhaled diesel exhaust and issued classifications. Some of these classifications occurred as long as two decades ago, and many have not been reviewed in light of the most recent information. In 1988, NIOSH declared diesel exhaust a “potential occupational carcinogen” (NIOSH, 1988). In 1989, the IARC classified diesel exhaust in Group 2A, “probably carcinogenic to humans, assessing the epidemiological evidence as limited and the animal evidence as sufficient” (IARC, 1989). In 1995, the HEI noted that, although the weight of epidemiological evidence suggested an association between occupational exposure to diesel exhaust and lung cancer, several uncertainties precluded judging the level of risk with confidence (HEI, 1995). In 1996, the World Health Organization’s International Programme on Chemical Safety (IPCS) classified diesel exhaust as “probably carcinogenic to humans” but noted that current

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epidemiological data were insufficient for estimating risk quantitatively (IPCS, 1996). The California EPA reviewed the carcinogenicity of diesel exhaust in view of its potential listing as a “toxic air contaminant” under state law. California considered the epidemiological data adequate to support a causal association between occupational exposures and lung cancer, and identified “diesel exhaust particulate matter,” in contrast to diesel exhaust per se, as a toxic air contaminant (Cal EPA, 1998). In 1998, the NTP reviewed “diesel exhaust particulates” for listing in its Report on Carcinogens (NTP, 1998). The Carcinogens Subcommittee of the NTP Board of Scientific Counselors voted to list diesel exhaust particulates as “reasonably anticipated to be a human carcinogen” but considered the evidence insufficient for the alternative listing as “known to be a human carcinogen.” In 2002, the U.S. EPA categorized diesel exhaust as “likely to be carcinogenic to humans” (EPA, 2002). In aggregate, the above designations portray a general consensus that diesel exhaust, and most probably DPM, poses some level of human lung cancer risk at some exposure level. There is also continuing concern for potential cancer risk among the general population from environmental exposures that overlap, in some locations such as those associated with heavy traffic, with occupational exposure levels. Most agencies and scientists agree that the present data do not allow estimation of unit cancer risks for humans with high confidence. It should be noted that all of the above assessments were based on results from studies to date; there are no data from either animals or humans from which to estimate carcinogenic hazards or risks from exhaust from the most recent fuel, engine, or after-treatment technologies. 16.4.2

Noncancer Health Effects

There has also been concern for noncancer health effects of diesel exhaust, both because of direct exposures and because of its contribution to general air pollution. These concerns include nonmalignant respiratory effects, such as chronic obstructive pulmonary disease, asthma, allergic sensitization, respiratory symptoms, and effects in other systems. Because DPM is ubiquitous and comprises a variable portion of ambient PM, it may contribute to these health risks in the general population, particularly in close proximity to traffic. Recent evidence suggests that several types of effects are at least plausible, and these topics will be reviewed here. 16.4.2.1 Amplification of Respiratory Allergic Responses It has been hypothesized that diesel emissions have contributed to the increased incidence of asthma and allergic rhinitis in developed countries over the past several decades (Pandya et al., 2002; Riedl and Diaz-Sanchez, 2005). In part, this concern is related to a broader question about links between air pollution and increased asthma; however, air pollution levels have fallen in most countries over the same period that asthma has increased. The specific role of diesel emissions was initially questioned in Japan, where a marked increase in allergy to cedar pollen following widespread planting of cedar trees during post-war reconstruction was accompanied by increased public exposure to diesel exhaust. Research in Japan, and later in the U.S., found that high exposures to diesel exhaust or DPM could enhance allergic responses, raising the hypothesis that DPM could act as an adjuvant. Although Wade and Newman (1993) used the term “diesel asthma,” the airway hyperreactivity and reversible airflow limitation they described in three railroad workers was very likely not due to allergic sensitization. The syndrome developed after extremely high exposures, suggesting a nonspecific reaction of the airways to irritants in high concentration. The findings

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reviewed below support the plausibility of a contribution of environmental exposures to diesel exhaust to respiratory allergies, but whether or not exhaust at environmentally relevant concentrations acts as an adjuvant in the development or exacerbation of allergies remains uncertain. Effects of Diesel Particles Research in Japan and the U.S. on the potential for DPM to amplify allergic responses was stimulated by the finding of Muranaka et al. (1986) that the IgE antibody response to ovalbumin injected intraperitoneally in mice was amplified by mixing DPM with the antigen. This study is also cited for the source of DPM used in several subsequent studies, including those described in the following paragraph. The DPM was generated by a 1980 Nissan automobile with an LD-28 2.97-L engine operated at speeds of 20–80 km/h on a chassis dynamometer. Particles were collected on filters from a constantvolume dilution tunnel. The PM composition was not reported, although a later paper (Tsien et al., 1997) indicated that the mass was 41% extractable by dichloromethane. Following this initial finding, Takafuji et al. (1987) elicited a similar amplification of allergic response in the respiratory tract by instilling ovalbumin and DPM simultaneously into the noses of mice (Takafuji et al., 1987). Maejima et al. (1997) compared the effects of different particles by dropping suspensions of DPM onto the nares of mice followed by inhalation exposure to Japanese cedar pollen weekly for several weeks. Kanto loam dust, DPM, carbon black, and coal fly ash all amplified antigen-specific IgE to a similar degree above the level in mice given pollen alone. Hao et al. (2003) sensitized BALB/c mice to ovalbumin by intraperitoneal injection with alum adjuvant, and then challenged them with aerosolized ovalbumin, either together with aerosolized DPM or followed by DPM. Treatment with DPM in either sequence enhanced the inflammatory response to antigen, but did not increase expression of antigen-specific antibodies. Researchers at the University of California Los Angeles (UCLA) have conducted numerous studies by instilling DPM intranasally in human subjects, typically at a dose of 300 mg. Japanese sources of DPM have been cited, but the composition of the material has not been reported. They first found that instillation of DPM increased IgE in nasal washings (Diaz-Sanchez et al., 1994). This was followed by demonstration that DPM also stimulated proinflammatory and proallergic cytokine production (Diaz-Sanchez et al., 1996). They found that combining DPM with ragweed allergen challenge markedly enhanced nasal ragweed-specific IgE and shifted cytokine production toward a Th2 pattern in ragweedsensitive subjects (Diaz-Sanchez et al., 1997). They inferred that soot-borne PAHs were the causal component class from finding that either a dichloromethane extract of DPM (incorrectly termed “PAH” by the authors) or pure phenanthrene increased IgE production in human B lymphocytes dosed in vitro (Takanaka et al., 1995; Tsien et al., 1997). They found that combined dosing with ragweed antigen and DPM caused IgE isotype switching (Fujieda et al., 1998). They found that pretreatment with topical fluticasone proprionate did not reduce the IgE or cytokine response to DPM, but did reduce the IgE and cytokine response to ragweed challenge (Diaz-Sanchez et al., 1999a). They instilled keyhole limpit hemocyanin (KLH), into na€ıve atopic subjects with or without pretreatment with DPM, and found that only the DPM-treated subjects developed KLH-specific IgE (Diaz-Sanchez et al., 1999b). This finding suggested that not only can DPM enhance responses in preallergic subjects, but it may also enhance development of allergies. Using a crossover study design, they found considerable individual consistency in the amplification by DPM of responses to ragweed challenge, and concluded that susceptibility to the adjuvant effect was an intrinsic trait (Bastain et al., 2003).

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In addition to the work described above, both the UCLA group and other investigators have reported numerous studies of the cellular mechanisms by which DPM might alter allergic responses. Most of these have been in vitro studies using epithelial or endothelial cells, lymphocyte subpopulations (e.g., Mamessier et al., 2006), or basophils. Rather than attempting a complete review, suffice it to note that enough mechanisms have been implicated to support the plausibility that, at some dose, DPM can initiate, alter, and amplify steps in allergic response pathways. Because other combustion-derived particles have not received the same scrutiny, the extent to which DPM might be unique in this respect is not known. Effects of Laboratory-Based Exposures to Diesel Exhaust Human Studies Investigators at the University of Umea in Sweden have conducted experimental exposures of normal and asthmatic subjects to exhaust from diesel vehicles operated at steady state. The compositions of the exposure atmospheres have not been described in detail; most exposures were described solely by DPM mass concentration. Exposure of atopic asthmatics for 1 h at 300 mg DPM/m3 increased airway resistance, airway hyperresponsiveness, and proinflammatory cytokines (Nordenhall et al., 2001). Exposures of normal and asthmatic subjects for 2 h at 100 mg DPM/m3 caused similar increases in airway resistance, but exposure did not enhance the eosinophilic inflammation in asthmatics, and the neutrophilic inflammatory response was less in asthmatics than in normal (Stenfors et al., 2004). Exposure of normal subjects for 1 h at 300 mg DPM/m3 caused an increase in IL-13 in bronchial mucosa, which is consistent with promotion of a TH2 inflammatory response (Pourazar et al., 2004). The ability of exposure to exacerbate allergic responses to antigen was not tested directly. However, the induction of inflammatory responses and proallergic cytokines in normal airways and increased reactivity of asthmatic airways indicate that acute high-level exposures can aggravate asthma and may promote allergic responses. Indirect evidence for a relationship between allergic airway disorders and inhaled diesel exhaust comes from studies of lung function in children in the Netherlands attending schools with different proximity to busy roadways. Brunekreef et al. (1997) reported a relationship between proximity of schools to busy roadways and reduced lung function, and found a more significant correlation with truck than with automobile traffic counts. Subsequently, Janssen et al. (2003) observed similar relationships with chronic respiratory symptoms, but found that the effect was almost exclusively expressed in children with bronchial hyperresponsiveness and/or sensitization to common allergens. These results do not confirm that diesel exhaust amplifies allergic responses, but they suggest that the effects of truck (nearly exclusively diesel) exhaust on lung function are expressed most strongly in children with asthma or allergies. Animal Studies Inhalation exposures of rodents to high concentrations of diesel exhaust have been shown to alter sensitization to antigens in some, but not all studies. Takano et al. (1998) exposed ICR mice 12 h/day, 7 days/week for 40 weeks to exhaust from a 2.7-L automobile engine operated at constant speed and load at DPM concentrations of 300, 1000, or 3000 mg/m3 (NOx ¼ 3.5, 10.1, and 23.1 ppm) or to clean air. The mice were sensitized to ovalbumin by intraperitoneal injection, and at 3-week intervals during the last 24 weeks, they were also exposed briefly to aerosols of ovalbumin. Exhaust exposure caused a concentration-related increase in neutrophils, eosinophils, and interleukin-5

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in bronchoalveolar lavage fluid, but only a very slight increase in serum antigen-specific IgE. Maejima et al. (2001) exposed female BDF1 mice 16 h/day, 5 days/week for 24 weeks to exhaust from a 6.9-L truck engine operated at constant speed and load at a DPM concentration of 3200 mg/m3 (NOx ¼ 30.4 ppm), diesel exhaust at the same dilution with DPM removed by filtration, Kanto loam dust at 3300 mg/m3, or clean air. On 2 days each week, Japanese cedar pollen was added to the exposure atmosphere at a concentration of 550,000 grains/m3. At the end of exposure, the group mean antigen-specific serum IgE levels were 4–5 fold higher in the diesel exhaust, filtered exhaust, and loam particle-exposed groups than in the pollen-only group. All three exposures significantly increased the percentages of mice having elevated titers. This finding indicated that either whole diesel exhaust or exhaust without DPM, and a mineral particle at the same concentration as DPM in whole exhaust, all amplified the sensitization of mice to pollen. A study by Watanabe and Ohsawa (2002) demonstrated the potential importance of exposures before and soon after birth in the later development of allergies. They exposed F344 rats 6 h/day for 19 days to exhaust from a 0.3-L single-cylinder engine operated at constant speed at a dilution yielding a DPM concentration of 1730 mg/m3 or to filtered exhaust at the same dilution. Exposures occurred over one of three periods: gestation day 7 to 3 days of age (in utero), 4–22 of age (postnatal), or 23–41 days of age (weanlings). At 49 days of age, all rats were sensitized to Japanese cedar pollen, and the resulting antigen-specific IgE titers were measured. The in utero and postnatal exposures increased IgE production to a similar degree, but the exposure after weaning did not. Filtered and unfiltered exhaust caused the same effect. No enhancement of allergic sensitization in mice was found by Barrett et al. (2002) using more environmentally relevant concentrations of exhaust. They exposed male BALB/c mice 6 h/day, 7 days/week for 8 weeks to exhaust from 5.9-L 2000 Cummins engines operated on U.S. certification fuel on the heavy-duty certification transient cycle at dilutions producing DPM concentrations of 30, 100, 300, and 1000 mg/m3 (NOx ¼ 2.2, 5.6, 16.9, and 49.3 ppm). The mice were also exposed at weeks 4 and 6 and after completion of exhaust exposure to aerosolized ovalbumin. Exposure to exhaust did not increase the sensitization of mice to ovalbumin, as measured by lung inflammation, proallergic cytokines, serum antigenspecific IgE, or airway responsiveness to methacholine. 16.4.2.2

Reduced Resistance to Respiratory Infection

Evidence from Humans There have been no direct studies of the impact of experimental exposures to diesel exhaust on the resistance of human subjects to respiratory infections. Present evidence is indirect, and derives from studies of relationships between air pollution or traffic and respiratory infections in children. In adults, robust associations between PM2.5 and hospital admissions for respiratory illnesses have been reported, including respiratory tract infections such as bronchitis and pneumonia obtained from Medicare claims data (Dominici et al., 2006). Kim et al. (2004) conducted a survey of approximately 1100 children in grades 3–5 in the San Francisco Bay area in 2001 whose schools were near a traffic corridor, and measured PM2.5, black carbon (as an index of combustion soot), and NO2 and calculated average exposure values. Adjusting for multiple other potential determinants of illness in children, such as a current smoker in the home and mold, there were small increases in the risk of bronchitis and asthma (2–4%

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increase per inter-quartile range of each pollutant). In a birth cohort of nearly 3000 children in the Netherlands (Brauer et al., 2002), an inter-quartile increase in soot concentration was associated with an increased risk of ear, nose, or throat infection (OR ¼ 1.15; 95% CI ¼ 1.00–1.33) and doctor-diagnosed flu or colds (OR ¼ 1.09; 95% CI ¼ 0.98–1.21), findings that approached statistical significance. The addresses of 11,484 patients from the National Cystic Fibrosis registry 1999–2000 were linked to the EPA Aerometric Information Retrieval Database (Goss et al., 2004). The risk of an exacerbation was significantly related to increases in PM2.5 but not NO2, adjusting for age, sex weight, and airway colonization with Pseudomonas. However, in a study of 19,901 infants discharged from a hospital after an episode of bronchiolitis in California during 1995–2000, there was no association with ambient CO, NO2, or PM2.5 (Karr et al., 2006). Taken together, these epidemiologic studies indicate that it is plausible that diesel exhaust exposure contributes to an increased susceptibility of respiratory infections in adults and children. Evidence from Animals Studies of rodents exposed to diesel exhaust by inhalation or DPM by intratracheal instillation have yielded mixed evidence for reducing resistance to bacterial respiratory infection, but generally support the plausibility that at least heavy exposures can do so. Campbell et al. (1981) exposed mice 8 h/day, 7 days/week to diesel exhaust at a dilution containing 6000 mg DPM/m3 and followed by challenge with Streptococcus pyogenes. Exposures ranging from 2 h to 321 days increased mortality from infection. Hatch et al. (1985) found no effect on mortality of intratracheal instillations of 100 mg DPM followed by S. pyogenes infection. Yang et al. (2001) found that intratracheal instillation of 5000 mg DPM/kg body weight into Sprague–Dawley rats depressed killing of Listeria monocytogenes instilled 3 days later, and that carbon black did not have the same effect. Harrod et al. (2004) exposed C57BL/6 mice 6 h/day for either 7 days or 6 months to exhaust from 5.9-L 2000 Cummins engines operated on U.S. certification fuel on the heavy-duty certification transient cycle at dilutions producing DPM concentrations of 30, 100, 300, and 1000 mg/m3 (NOx ¼ 2.2, 5.6, 16.9, and 49.3 ppm), followed by intratracheal instillation of Pseudomonas aeruginosa. Although clearance of bacteria was slowed and infection-related histopathology was increased in a generally concentration-dependent manner at both exposure times, the effect was only significant for the 7-day exposure. Experimental exposures have also produced mixed results for effects on resistance to respiratory viral infection. Hahon et al. (1985; also described in Castranova et al., 2001) exposed CD-1 mice 7 h/day 5 days/week for 1, 3, or 6 months to exhaust at a DPM concentration of 2000 mg/m3 followed by intranasal instillation of Ao/PR/8/34 influenza virus. Although mortality was not affected by exposure, exposure for 3 or 6 months increased viral replication in the lung and lung histopathology and suppressed the production of interferon. Harrod et al. (2003) exposed C57BL/6 mice for 7 days to exhaust as described in the preceding paragraph, but only at the 30 and 1000 mg DPM/m3 levels, followed by intratracheal instillation of Respiratory Syncytial Virus (RSV). In that pilot study, viral clearance was reduced, and infection-related histopathology was increased at both exposure levels. In a follow-up study using the same exposure but all four exposure levels, Reed and Berger (2006b) found no effect on viral clearance or histopathology at any level. The reason that the effects in the pilot study were not reproduced later was not confirmed, but experience has shown that the assay produces variable results depending on the specific viral culture used and the viral titering methods (both differed between the studies).

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Potential Mechanisms The mechanisms by which exposure to diesel exhaust might reduce resistance to respiratory infections are not well-defined, but two factors are suggested by previous work: suppression of macrophage function and suppression of systemic immune responses. Macrophage cytotoxicity has been associated with oxidative stress produced by DPM extracts (Hiura et al., 1999; Li et al., 2000), and it has been postulated that quinones might be an important chemical class (Li et al., 2000; Kumagai et al., 2002). Although allergic responses might be amplified by exposure to diesel exhaust as discussed above, protective immune responses might be suppressed. Evidence that diesel exhaust can suppress systemic immune responses is discussed in a following section. 16.4.2.3 Respiratory Tract Inflammation and Noncancer Respiratory Disease Studies of occupational groups having high exposure to diesel exhaust show acute effects on lung function from high exposures and higher incidences of noncancer lung disorders. Several studies of experimentally exposed humans have shown that acute exposures to concentrations within the high occupational range can induce inflammation, reduce lung function, and increase airway reactivity. Short- and long-term exposures of animals at high concentrations produce inflammatory responses. Chronic, extreme exposures of rats produce chronic active lung inflammation, an overwhelming of particle clearance, progressive particle sequestration, fibrosis, and epithelial hyperplasia, metaplasia, and neoplasia. Effects in Humans Experimental Exposures Battigelli (1965) exposed subjects for 1 h to exhaust from a single-cylinder diesel engine and detected no decrements of airflow resistance. Although the DPM concentrations were not given, the highest concentrations of 55 ppm CO, 4.2 ppm NO2, and 1 ppm SO2 suggest that the highest soot concentration was in the range of a few mg/m3. Ulfvarson et al. (1987) exposed subjects for 3.7 h to exhaust from a 3.7-L engine at 600 mg DPM/m3 and detected no decrement in pulmonary function. In a series of experiments conducted by Swedish investigators, Rudell et al. (1996, 1999 exposed healthy, never-smoking subjects to diesel exhaust from an idling truck for 1 h. PM mass concentration was not specified, but eye irritation, nasal irritation, and an unpleasant smell was reported during exposure. Although airway resistance increased, there was no significant difference in FVC, FEV1, or other flows, but there was an increase in neutrophils and decreased phagocytosis by macrophages in bronchoalveolar lavage fluid. In additional studies by the same Swedish group, healthy subjects were exposed to diesel exhaust at 300 mg DPM/m3 for 1 h while riding a stationary bicycle. As before, there was no change in spirometry, but there was an increase in neutrophils and B-lymphocytes in bronchoalveolar lavage fluid. There was also an increase in neutrophils; mast cells; CD4þ, and CD8þ T lymphocytes; with upregulation of vascular endothelial cell adhesion molecules and in bronchial biopsies and evidence of increased IL-8 gene transcription (Salvi et al., 1999, 2000). In peripheral blood sampled 6 h after exposure, there was a significant increase in neutrophils and platelets (Salvi et al., 1999), and also greater numbers of neutrophils in expectorated sputum and IL-6 compared to subjects exposed to air (Nordenhall et al., 2000). In a similar study in Britain, subjects were exposed to either air or diesel exhaust at 200 mg DPM/m3 for 2 h while at rest (Nightingale et al., 2000). There were no changes in spirometry, but there was an increase in expectorated sputum neutrophils and myeloperoxidase 4 h after exposure with no changes in peripheral blood neutrophils. More recently, the Swedish group

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exposed volunteers to diesel exhaust at 100 mg DPM/m3 for 2 h, and at 18 h post-exposure bronchoalveolar lavage and bronchial biopsies were obtained (Behndig et al., 2006). Increased bronchial mucosa neutrophils, mast cells, and increased neutrophils, IL-8, and myeloperoxidase were noted in the lavage fluid. These findings indicate that exposure of healthy subjects to diesel exhaust at levels of 100 mg DPM/m3 can cause a pulmonary inflammatory response, influence macrophage function, and result in the expression of airway cytokines, without causing significant changes in ventilatory function measured using standard clinical tests. There may also be changes in vascular function and inflammatory changes in peripheral blood, suggesting a link to systemic effects occurring after diesel exhaust inhalation. Effects from a Single Workshift There are six published evaluations of the effects of diesel exhaust exposure on respiratory function during a single workshift. These studies are complicated by the numerous other materials inhaled during the workshift by the miners, garage workers, stevedores, and ferryboat crewmen, and not all studies included control groups not exposed to diesel exhaust. In addition, characterization of the personal exposures to diesel soot and other exhaust components varied considerably among the studies. All of these studies used spirometry to assess respiratory function. Ames et al. (1982) found that workshift changes in respiratory function did not differ between workers in mines where diesel engines were used or not used. Both J€ orgensen and Svensson (1970) and Gamble et al. (1978) reported small workshift decrements in forced expiratory volumes and flow rates among miners. Gamble et al. (1987) measured small workshift decrements in function, and found a stronger association with PM than with NO2. Ulfvarson et al. (1987) found small, but significant workshift decrements in function among stevedores on roll-on, roll-off ships exposed to DPM at 130–590 mg/m3, but not in bus garage or car ferry workers exposed to DPM at 100–460 mg/m3. Ulfvarson and Alexandersson (1990) detected small workshift decrements of function among stevedores exposed to PM at 120 mg/m3. Overall, these findings support the conclusion that reversible changes in respiratory function can occur in humans exposed occupationally, although it is not possible to relate these changes to a specific level of exposure. Long-Term Effects on Respiratory Function and Symptoms Evaluations of the effects of longer-term occupational exposures to diesel exhaust on respiratory function and symptoms have yielded mixed results. Most studies found that exposure was associated with small increases in respiratory symptoms, such as dyspnea, cough, and phlegm (J€orgensen and Svensson, 1970; Attfield et al., 1982; Reger et al., 1982; Gamble et al., 1983; Purdam et al., 1987), but some did not (Battigelli et al., 1964; Ames et al., 1984). There was no consistent effect on respiratory function, but it is possible that the lack of a clear relationship was due to including few workers with long-term exposures and by only including active workers in these studies. In a more recent study, respiratory symptoms in 20,898 farmers were assessed in relation to occupational exposures (Hoppin et al., 2004). Adjusting for age, smoking, asthma history, and atopy, driving diesel tractors was associated with an elevated OR for wheeze (1.31; 95% CI ¼ 1.13–1.52), and there was an increasing risk with increasing years of tractor driving. For farmers driving gasoline tractors, the overall risk was also significantly elevated but of lower magnitude (1.11; 95% CI ¼ 1.02–1.21). Although these results support an association between diesel exposure and respiratory symptoms, the interpretation is hampered by lack of actual exposure information. Jacobson et al. (1988) found in a cohort of 19,901 British coal miners investigated over a 5-year period

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that there was an increased work absence due to self-reported chest illness in underground workers exposed to diesel exhaust, as compared to surface workers. This finding is consistent with the impairment of alveolar macrophage function or may be due to the effects of exposure on airway inflammation noted following acute exposure described in the previous sections. Traffic Studies There is a large body of emerging literature, primarily in children, relating proximity to traffic to chronic respiratory symptoms. It has not been possible to exclusively implicate diesel exhaust in the occurrence of symptoms. Venn and coworkers (2001) reported that the risk of wheeze among children living in Nottingham, England, within 150 meters of a main road increased 8% for every 30 meters closer in distance, with most of the risk within 90 meters of the road. In the Netherlands, a doubling of the risk of wheeze was reported in children living within 100 meters of a roadway (Van Vliet et al., 1997). For children living within 300 meters, measurements of truck traffic density (mostly diesel) but not automobile traffic density was significantly related to lower values of pulmonary function (Brunekreef et al., 1997). Self-reports of truck traffic volume were also significantly related to wheezing (Weiland et al., 1994; Duhme et al., 1996) and recurrent respiratory illness (Ciccone et al., 1998). An association between ambient NOx, whose major source is diesel vehicle emissions, and asthma prevalence among Taiwan middle-school students was also noted (Guo et al., 1999). The International Study of Asthma and Allergies in Childhood study in Munich Germany assessed nearby residential traffic counts for 7509 children (Nicolai et al., 2003). Current asthma, wheeze, and cough were significantly related to traffic counts. There are fewer traffic studies for adults. Oosterlee et al. (1996) found no association with chronic cough, chronic phlegm, and wheeze in adults living on busy streets compared to those living in neighborhoods with little traffic. In Tokyo, three cross-sectional studies were conducted in three separate groups of over 1500 women in 1979, 1982, and 1983, and conflicting results were obtained (Nitta et al., 1993). Garshick et al. (2003) studied male U.S. veterans drawn from the general population of southeastern Massachusetts. Information on respiratory symptoms and potential risk factors were collected by questionnaire, and residential addresses were related to distance from and traffic density of major roadways. Adjusting for cigarette smoking, age, and occupational exposure to dust, subjects living within 50 meters from a major roadway were more likely to report persistent wheeze (OR ¼ 1.31; 95% CI ¼ 1.00–1.71) compared to those >400 meters away. The risk was only observed for those living within 50 meters of heavily trafficked roads (10,000 vehicles/ 24 h): OR ¼ 1.71, 95% CI ¼ 1.22–2.40). The risk of chronic phlegm on heavily trafficked roads was of borderline significance (OR ¼ 1.40, 95% CI ¼ 0.97–2.02). These results suggest that residential exposure to vehicular emissions near busy roadways (to which diesel emissions contribute, but not exclusively) results in chronic respiratory disease symptoms in adults and children and may be associated with asthma. As evidence that a reduction in traffic results in an improvement in asthma morbidity, during the Atlanta Olympic games in 1996 when traffic was restricted, there were lower rates of childhood asthma hospitalizations, Medicaid claims, and emergency room utilization (Friedman et al., 2001). Nonmalignant Respiratory Disease Mortality There have been few studies by which the relationship between long-term occupational exposure to diesel exhaust and mortality from noncancer respiratory disease can be assessed (HEI, 1995). Boffetta et al. (1988) not only described lung cancer mortality in the American Cancer Society Cohort as described

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earlier, but also examined mortality attributable to respiratory diseases. Adjusting for smoking and age, chronic obstructive pulmonary disease (COPD) was not significantly increased, but mortality attributable to pneumonia and influenza was significantly increased (p < 0.05; relative risk ¼ 1.97). Wong et al. (1985) assessed mortality in 34,156 operating engineer union members during 1964–1978, and a greater emphysema mortality risk was associated with a greater duration of union membership. In the 38-year mortality followup of U.S. railroad workers, Garshick et al. (2004) observed an elevated mortality risk from all respiratory system diseases, including COPD and allied conditions (relative risk ¼ 1.31; 95% CI ¼ 1.21–1.42), and from COPD and allied conditions alone (1.41; 95% CI ¼ 1.27–1.55. Because there was no direct information regarding smoking, the same authors conducted a case–control study using deaths collected during 1981–1982 with smoking histories from next of kin (Garshick et al., 1987; Hart et al., 2006). In that study there were 536 cases with COPD or allied conditions and 1525 controls with causes of death not related to diesel exhaust or fine particle exposure. After adjustment for age, race, smoking, U.S. Census region of death, and vitamin C use, engineers and conductors (occupations with diesel exposure from operating trains) had an increased risk of COPD mortality. Mortality increased with years of work in jobs with exposure, and for 16þ years of exposure after 1959, OR ¼ 1.61 (95% confidence interval ¼ 1.12–2.30). Overall, the risk for mortality from noncancer chronic respiratory disease associated with occupational exposure to diesel exhaust appears to be on the same order of magnitude as the risk for lung cancer, but the data supporting the association are fewer. Effects in Animals Repeated exposure of animals to diesel exhaust induces concentration-related effects on respiratory function and lung structure that have been reviewed previously (Mauderly, 1994a, 1996, 2000; HEI, 1995; EPA, 2002). Recent studies provide additional detail (e.g., Kato et al., 2000), but are consistent with earlier findings. Near lifetime repeated exposures of rats at concentrations of DPM over approximately 1000 mg/m3 overwhelms the ability of normal particle clearance pathways and results in a progressive accumulation of DPM in the lung. This accumulation is accompanied by persistent inflammation, focal epithelial proliferation and metaplasia, and fibrosis (Mauderly, 1996). The progressive structural changes are reflected by a progressive impairment of respiratory function that includes lung stiffening (loss of compliance), reduced lung volumes, uneven intrapulmonary gas distribution, and impaired alveolar-capillary gas exchange (Mauderly et al., 1988). This structure-function syndrome also occurs in rats exposed heavily to other solid, respirable particles (Mauderly, 1994a). Of importance for estimating hazard for humans, no significant alterations of particle clearance (Wolff et al., 1987), inflammation, fibrosis (Henderson et al., 1988) or respiratory function or structure (Mauderly et al., 1988) resulted from chronic exposures of rats at 350 mg DPM/m3, even though small amounts of DPM accumulated in the lungs. This dose-response information from rats was used by EPA to estimate a reference (safe) concentration for noncancer effects in humans of 5 mg DPM/m3 lifetime exposure (EPA, 2002), after adjustments for interspecies dosimetry and safety factors. A study by Kato et al. (2000) was consistent with the earlier findings in detecting no significant tissue effects of exposure of rats for 24 months at 210 mg DPM/m3. Under exposure conditions producing the above effects in rats, mice accumulate similar amounts of DPM in their lungs (Henderson et al., 1988), but the inflammatory, fibrotic (Henderson et al., 1988), and histopathological (Mauderly et al., 1996) responses are less than those in rats. Small reductions in lung volumes and compliance have also been observed

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in diesel-exposed Syrian (Heinrich et al., 1986) and Chinese (Vinegar et al., 1981) hamsters. Mice and hamsters have also been shown to have lesser functional and structural responses than rats to other solid respirable particles (Mauderly, 1994b). A smaller, but perhaps more relevant, body of information comes from nonrodent species that have respiratory bronchioles (absent in rodents) and other lung features more similar to humans. Only two such species have been chronically exposed to diesel exhaust. The EPA exposed male cats chronically to exhaust at 6000 mg DPM/m3 for 61 weeks, and then at 12,000 mg DPM/m3 for the remainder of 27 months, followed by a 6-month recovery period (Pepelko and Peirano, 1983). A restrictive functional impairment with decreased lung volumes and uneven intrapulmonary gas distribution was observed at the end of the exposure (Moorman et al., 1985). Histopathology at the end of exposure included peribronchiolar fibrosis and epithelial metaplasia in terminal and respiratory bronchioles (Plopper et al., 1983). Interestingly, while the epithelial changes lessened during the 6-month recovery period, the fibrosis progressed. Lewis et al. (1989) exposed cynomolgus monkeys to diesel exhaust at 2000 mg DPM/m3 and reported that the forced expiratory flow rates were reduced at the end of exposure (Lewis et al., 1986). The lung histopathology of the monkeys differed from that of rats exposed concurrently (Nikula et al., 1997). Soot was present in approximately the same tissue concentration in both species, but was located predominantly in interstitial compartments in monkeys and in alveolar lumens in rats. The species had similar increases in pulmonary macrophages. The most striking difference was in the degree of epithelial proliferation, which was characteristically prevalent near accumulations of soot in rats but essentially absent in monkeys. Although the data base is small, these results suggest that nonrodent species can develop fibrosis and epithelial responses under extreme exposure conditions but exhibit little structural response from chronic exposures at 2000 mg DPM/m3. 16.4.2.4 Cardiovascular Effects There has been increasing attention to the potential cardiovascular effects of inhaled diesel exhaust, concurrent with increasing evidence for the cardiovascular effects of ambient PM (EPA, 2004). Similar to the effects described for other organ systems, it is now clear that exposures to high concentrations of exhaust have potential for altering heart and vascular function in both humans and animals, but the hazards and risks from typical environmental exposures are not yet known. The plausibility of an effect from environmental exposures to DPM is supported by evidence for the effects of ambient PM2.5, to which DPM contribute. In 1995, Seaton and coworkers suggested that lung inflammation caused by inhaled ultrafine PM in might provoke myocardial infarction (MI) as a result of mediator release (Seaton et al., 1995). In recent years it has been recognized that atherosclerosis and coronary artery disease are chronic inflammatory diseases and that blood markers of systemic inflammation are related to clinical cardiovascular outcomes, including MI (Ross, 1999; Pearson et al., 2003). Elevated blood levels of cardiovascular inflammatory markers such as C-reactive protein are associated with increases in particulate air pollution (Schwartz, 2001; Peters et al., 2001; van Eeden et al., 2001). Laboratory studies also suggest a link between PM2.5 and vascular changes, including atherosclerosis. Suwa et al. (2002) instilled urban PM from Ottawa into the nasopharynx of rabbits genetically prone to atherosclerosis and observed a progression of atherosclerotic lesions accompanied by an increase in circulating neutrophils. Sun et al. (2005) observed an enhancement of atherosclerotic changes in the aortas of ApoE / mice exposed to concentrated Northeastern regional PM2.5 6 h/day, 5 days/week for 6 months.

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Other studies of humans and animals have shown associations between ambient PM and heart rate control, arrhythmias, and abnormalities of cardiac depolarization as manifested by ST segment depression. Although none of the ambient PM results are specific to DPM, DPM must be assumed to contribute some portion in lieu of better information on the PM species causing the effects. Studies in Humans Information from humans derives from several study designs. There is indirect evidence from occupational and population-based epidemiologic studies linking adverse cardiovascular outcomes to traffic emissions, evidence from panel studies where smaller groups were monitored and physiologic observations were recorded and related to environmental exposures, and direct evidence from laboratory studies involving experimental exposures. Experimental and Panel Studies of Traffic, Ambient PM, and Black Carbon There is evidence that inhaled ultrafine carbon particles, similar to the EC core of diesel soot, can pass through the lung into the systemic circulation in humans. Nemmar et al. (2002) administered 99m Technetium-labeled carbon particles via inhalation to five healthy volunteers. There was a rapid increase in radioactivity over the liver, and radioactivity counts increased in blood samples and urine. Pekkanen et al. (2002) assessed ST segment depression during exercise in 342 tests of 45 subjects with coronary disease in Helsinki during the winters of 1998–1999. Air pollution was monitored at a central site, and levels of particulate air pollution 2 days before the test were associated with an increased risk of ST segment depression. The central site recorded the number of ultrafine particles (10–100 nm), number of particles 100–1000 nm (accumulation mode), and particle mass as PM2.5. Each PM measure was associated with an increased risk of ST-segment depression, but the effects of ultrafine PM and PM2.5 were independent, suggesting separate sources. In a subsequent analysis of the same data, filter absorbance was assessed as a measure of EC or black carbon (Lanki et al., 2006). Filter samples were also analyzed for elemental composition using X-ray fluorescence spectrometry and principal component analysis to apportion PM mass to sources. Traffic- and diesel-related PM as indicated by the extent of filter absorbance were associated with the greatest risk of ST depression (Lanki et al., 2006). Gold et al. (2005) studied ST segment changes in 24 active Boston residents 61–88 years of age, each monitored up to 12 times from June through September 1999. Black carbon level in the previous 12 h and the level 5 h before testing predicted ST-segment depression. These results suggest an adverse effect of traffic-related emissions, including diesel exhaust, on people likely to be at risk for cardiac ischemic events. Peters and coworkers (2004) in Augsburg, Southern Germany, found a significant association between exposure to traffic and the onset of a MI 1 h later (OR ¼ 2.92; 95% CI ¼ 2.22–3.83). Because Europe has large numbers of light-duty diesel vehicles, it was speculated that diesel exhaust significantly contributed to traffic emissions in that study. In an additional study in Erfurt, Germany, ambient PM was measured continuously during the winter of 2000–2001 (Ruckerl et al., 2006). Ultrafine and accumulation mode PM were both significantly related to increased levels of C-reactive protein (CRP) with a 2-day lag in 57 male patients with coronary heart disease. Associations between CRP and EC were weaker, but EC was strongly associated with increased levels of intercellular adhesion molecule-1, an indicator of vascular endothelial cell activation. In a study of nine North Carolina State troopers observed over 4 days, PM2.5 measured inside the patrol cars was associated with

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CRP measured post-shift, further suggesting an association between traffic-related exposures and biomarkers of cardiovascular disease (Riediker et al., 2004). There have also been a series of studies assessing the relationship between ambient air pollution and ventricular and supraventricular ectopy. The results indicate positive associations between ectopy and air pollution, but results based on specific pollutants have varied, and associations with black carbon as an index of traffic or diesel exposure have been inconsistent. Sarnat et al. (2006) assessed 32 nonsmoking older adults on a weekly basis for 24 weeks during the summer and autumn of 2000 in Steubenville, Ohio, using a standardized 30-min protocol that included continuous electrocardiogram measurements. There were associations between supraventricular ectopy and PM2.5, sulfate, and ozone concentrations but not with EC, and there were no associations between any pollutant and ventricular ectopy. In a study of 56 people with implantable defibrillators in St Louis (Rich et al., 2006a), a significant increase in risk of ventricular arrhythmias was associated with SO2 and an insignificantly increased risk was associated with increases in NO2 and EC in the 24 h before the arrhythmia. In a similar study in Boston, Rich et al. (2006b) found a statistically significant positive association between episodes of paroxysmal atrial fibrillation and increased ozone concentration in the hour before the arrhythmia, and positive but not statistically significant risks were associated with PM2.5, NO2, and black carbon. In another study in Boston (Rich et al., 2005), PM2.5 and ozone were associated increased risks of ventricular arrhythmia, and Dockery et al. (2005) found an increased risk of ventricular arrhythmia associated with a 2-day mean exposure for PM2.5, CO, NO2, and black carbon that were not statistically significant. However, statistically significant associations were noted in people who had an arrhythmia 3 days earlier. Peters et al. (2000) found associations with defibrillator discharges in Boston between ventricular arrhythmias and NO2. In a study assessing heart rate variability, a measure of autonomic heart rate regulation in elderly Boston residents, Schwartz et al. (2005) found the strongest associations between decreased variability and black carbon. Taken together, these results indicate an association between arrhythmias with air pollution that is suggestive of traffic-related sources. Laboratory Exposures to Diesel Emissions Mills et al. (2005) exposed 30 healthy men to diesel exhaust at 300 mg DPM/m3 or clean air while riding a bicycle ergometer for 1 h in a double-blind manner. Exhaust was generated by an idling 1991 model off-road engine. Forearm blood flow was monitored, and the endothelium-dependent vasodilators bradykinin and acetylcholine, and the endothelium-independent vasodilator nitropusside were administered 2–4 h after exposure. The vasodilators caused an increase in forearm blood flow that was blunted following exposure, suggesting that diesel exhaust inhibited regulation of vascular tone. Exposure also suppressed the increase in plasma tissue plasminogen activator following infusion of bradykinin. Because tissue plasminogen activator is an endogenous fibrinolytic agent, it was postulated that impaired release could impair the ability of diesel-exposed people to dissolve endogenous intravascular thrombi. Impaired endogenous fibrinolysis and impaired vascular tone could contribute to a greater risk for cardiovascular events. The group also repeated measurements of forearm blood flow and plasma markers at 24 h after exposure (T€ ornqvist et al., 2007). At that time, endotheliumdependent vasodilation was still reduced, but endothelium-independent vasodilation was not. Increases in plasma cytokines indicated that mild systemic inflammation persisted to 24 h.

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Mills et al. (2007) followed the above studies with an evaluation of effects on men with prior myocardial infarction. Twenty men with stable coronary artery disease underwent the same exposure protocol used in the preceding studies. Myocardial ischemia was evaluated by analysis of the ST segment of the electrocardiogram during exposure; plasma markers and responses to vasodilators were measured at 6 h after exposure, using the previous methods. A greater exercise-induced ST segment depression after diesel exposure than after sham exposure indicated a proischemic effect. The responses to the vasodilators were not altered by exposure, in contrast to the above findings with healthy men. The release of endothelial plasminogen activator was reduced by exposure. Together, the results from this group indicated that exposure to a high concentration of diesel emissions can cause acute cardiovascular changes, some of which may persist for 24 h. Occupational Studies There have been a number of studies examining cardiac outcomes in people exposed occupationally to diesel exhaust and other traffic-related pollutants. Taken as a whole, these studies indicate that professional drivers are at greater risk for cardiovascular disease than other occupations. The magnitude of the contribution of traffic emissions is uncertain because these studies have lacked exposure measurements. Tuchsen and Endahl (1999) described an excess of ischemic heart disease in Danish bus drivers. Employed men in Denmark aged 20–59 years old in 1981, 1986, and 1991 were classified according to occupation, and followed from 1981 to 1985, 1986 to 1990, and 1991 to 1993. All hospitalizations were identified using national records and ischemic heart disease rates in bus drivers were compared to rates for all employed men. In 1981–1985, the ischemic heart disease admission risk for bus drivers was significantly increased (standardized hospital admission ratio ¼ 1.41; 95% CI ¼ 1.20–1.65), and the risk increased throughout the study period. Hannerz and Tuchsen (2001) subsequently examined hospital admission rates for all Danish male professional drivers through 1997 and reported elevated ischemic heart disease rates for both taxi and bus drivers. Bigert et al. (2004) also conducted a study examining trends in the incidence of MI among professional drivers (1183 cases and 6072 controls) in Stockholm County, Sweden, during 1977–1996. They used registers of hospital discharges and deaths, and controls selected randomly from the general population (20,364 cases and 136,342 controls) that were linked to national occupational register job titles. Between 1977 and 1984, the risk among bus, taxi, and truck drivers was increased compared to manual workers, but risks subsequently declined among all three through 1996. Finkelstein et al. (2004) assessed mortality due to ischemic heart disease (1416 cases, ICD9 codes 410–414) in seven Ontario construction unions, and compared deaths among heavy equipment operators to deaths in other occupations. The construction workers had an increased risk of ischemic heart disease mortality (OR ¼ 1.32; 95% CI ¼ 1.13–1.55). Although there was no information regarding smoking or other risk factors in these studies, the comparison groups were other workers, making it unlikely that cardiovascular risk factors substantially differed among case and referent groups. The majority of other epidemiologic studies describing an association between cardiovascular disease and professional drivers were conducted in Sweden. Edling et al. (1987) followed 694 men in five Swedish bus companies during 1951–1983. Standardized mortality ratios were computed for clerks, bus drivers, and bus garage workers, and no increase in mortality was noted for cardiovascular diseases; however, there were relatively few deaths.

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In contrast, Gustavsson and coworkers (1996) studied cases of MI in professional drivers aged 30–74 in Stockholm County (4105 cases and 13,066 referents) during 1976–1984. Incident cases of first MI were identified using a central register. Referents were selected from the general population and matched based on age, year of admission, and death. National census data on occupation and industry was used to identify long-term drivers. The relative risk of MI was 1.53 (95% CI ¼ 1.15–2.05) for bus drivers (70 cases), 1.65 (95% CI ¼ 1.30–2.11) for taxi drivers (101 cases), and 1.31 (95% CI ¼ 1.05–1.64) for longdistance truck drivers (112 cases). Gustavsson et al. (2001) addressed specific occupational and personal risk factors for MI as part of the Stockholm Heart Epidemiology Program during 1992–1993 for men and during 1992–1994 for women in living in Stockholm County, and free of previous MI (1335 cases and 1658 controls). Cases surviving at least 28 days post infarction were identified using hospital data, and referents were identified through a computerized population register and matched on age, year of enrollment, and hospital service area. Participants completed mail questionnaires to obtain lifetime occupational histories and other personal life-style factors, with followup by telephone. Exposures were assessed using a job-exposure matrix designed to assess the intensity of exposure to motor exhaust using estimated CO levels found in various occupational categories. After adjusting for smoking, alcohol, body mass index, diabetes, hypertension, and physical activity, there was an increase in risk of MI for participants having medium exhaust cumulative exposure levels (154 cases and 136 controls; RR ¼ 1.32; 95% CI ¼ 1.01–1.73) and with high cumulative exhaust exposures (155 cases and 137 controls; RR ¼ 1.21; 95% CI ¼ 0.91–1.59). These data were extended by Bigert et al. (2003) who assessed MI risks among male bus, truck, and taxi drivers who had their first MI in 1992 or 1993. There were 1067 cases and 1482 controls, including 146 cases and 129 controls who had ever worked in a driver category. Adjusting for smoking, alcohol, physical activity during leisure time, body mass index (BMI), diabetes, hypertension, socioeconomic status, and an index of job stress, the MI risk among truck drivers (45 cases and 31 controls) was not elevated (OR ¼ 1.07; 95% CI ¼ 0.77–1.50). For bus drivers (45 cases and 31 controls; OR ¼ 1.46; 95% CI ¼ 0.89– 2.41) and taxi drivers (94 cases and 84 controls; OR ¼ 1.32; 95% CI ¼ 0.81–1.50) risk estimates were elevated (although not significantly so) with suggestion of an increased risk with greater years of work. Epidemiological Studies of Ambient PM There are associations between PM2.5 and short-term hospital admissions for cerebrovascular disease, peripheral vascular disease, ischemic heart disease, heart failure, and heart rhythm disturbances using Medicare claims data (Dominici et al., 2006). Pope et al. (2004) assessed cardiovascular mortality over a 16year period in the American Cancer Society cohort, and linked residential address to PM2.5 data for U.S. metropolitan areas for over 300,000 people. Long-term PM exposures were significantly associated with mortality attributable to ischemic heart disease, dysrhythmias, heart failure, and cardiac arrest. A followup of the Harvard Six City study assessed mortality during 1974–1998, and also found an association between cardiovascular mortality and average PM2.5 (Laden et al., 2006a). Laden et al. (2000) also examined daily mortality in these cities during 1979–1988 and found it to be significantly related to mobile source emissions using particle lead as a marker of exposure to gasoline vehicle emissions. Kunzli et al. (2005) used baseline health data from two studies assessing carotid intima-media thickness in 798 people in the Los Angeles area. Residential addresses were geocoded

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to assign annual mean PM2.5 values using a regional exposure model. There was a significant relationship between carotid intima-media thickness and PM2.5, adjusting for age, sex, smoking, blood pressure, and treatment for hypertension or with lipid lowering agents, thereby supporting an association between a direct measure of atherosclerosis and PM2.5. Although not specifically implicating diesel exhaust, these studies support a relationship between cardiovascular disease and PM2.5 to which diesels contribute. Other air pollution studies have included exposure markers that more specifically assessed a potential contribution from diesel exhaust exposure. London daily air pollution data were analyzed in relation to daily deaths during 1992–1994 (Bremner et al., 1999). Black smoke with a 1-day lag was a significant predictor of all cardiovascular mortality. Le Tertre et al. (2002) studied the relationship between airborne particles and hospital admissions for cardiac causes in eight European cities. These data included black smoke for in five cities (Barcelona, 1994–1996; Birmimgham, 1992–1995; London, 1992–1994; Netherlands, 1989–1995; Paris, 1992–1996). Black smoke with 0- and 1-day lags was associated with a 1.1% increase in cardiac deaths (95% CI ¼ 0.4–1.8) that was independent of total PM exposure. Hoek et al. (2002) studied a sample of 5000 people from the Netherlands Cohort Study on Diet and Cancer during 1986–1994. Cardiopulmonary mortality was associated with living near a major road (RR ¼ 1.95; 95% CI ¼ 1.09–3.52), defined as 100 m of a freeway or within 50 m of a major urban road, adjusting for smoking, education, BMI, and diet. Nafstad et al. (2004) found a significant association between estimated residential NO2 levels and ischemic heart disease in a population-based prospective study in Oslo, Norway, adjusting for smoking and other potential confounding factors. Rosenlund et al. (2006) conducted analyses using data from the Stockholm Heart Epidemiology Program. Addresses were geocoded and emission databases used to obtain estimates of annual mean levels of NO2, CO, PM10, PM2.5, and SO2. There was no overall association between long-term air pollution exposures and overall MI incidence, including NO2 exposures as an index of traffic-related exposures. Results were adjusted for age, catchment area, smoking, physical activity, diabetes, BMI, and occupational exposures. However, there was a suggestion of an increase in fatal cases attributable to NO2, (272 cases; OR ¼ 1.51; 95% CI ¼ 0.96–2.37), and risk was particularly elevated in those who died out of the hospital (84 cases; OR ¼ 2.17; 95% CI ¼ 1.05–4.51). Studies in Animals and In Vitro Systems Ikeda et al. (1995) found that incubating sections of rat aortas with suspensions of 1–100 mg DPM/mL in saline reduced acetylcholine-induced relaxation and that superoxide dismutase partially abolished the effect. They interpreted the results as suggesting that the DPM inhibited the production or effect of endothelial nitric oxide. Toda et al. (2001) injected 12,000 or 120,000 mg DPM/kg intravenously into rats and monitored blood pressure. The higher dose caused a transient drop in blood pressure that was blocked by atropine but not propanol. These results suggest that the small amounts of DPM that might circulate in the blood at any given time would not be expected to affect blood pressure in rats. Campen et al. (2003) found that repeated exposures altered the electrocardiograms of genetically hypertensive rats. Rats were monitored by telemetry before, during, and after exposure to diesel exhaust at 30, 100, 300, and 1000 mg DPM/m3 for 6 h/day for 7 days (exposure described in detail in McDonald et al., 2004a, and Figs. 16.1, 16.2, 16.3, and

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16.4). The mean heart rate was increased during the exposure period, largely because the heart rate of exposed rats did not tend to decline during the study as in controls and because the normal diurnal reduction during daytime was less in exposed rats. The effect was clear at even the lowest exposure level in males, but not females. The PQ interval was increased slightly by exposure, but significantly at the highest levels. This effect was also greater in males. Campen et al. (2005) demonstrated that the non-PM components of diesel exhaust had effects on the electrocardiogram by exposing C57BL/6 mice or atherosclerosis-prone ApoE / mice (pre-fed high fat diet) by inhalation. The mice were exposed 6 h/day for 3 days to whole exhaust from a single-cylinder engine operated under steady state load on certification fuel (McDonald et al., 2004c) at dilutions containing 500 or 3600 mg DPM/m3 or to filtered exhaust at the same dilutions. Both levels of whole and filtered exhausts caused identical concentration-related bradycardia in ApoE mice. Significant reductions of the T-wave area were caused by both whole and filtered exhaust at the high, but not low, concentration in ApoE mice, but not in C576BL/6 mice. They also perfused isolated coronary arteries from ApoE mice in vitro with saline through which the exhaust had been bubbled and found increased flow resistance and decreased response to vasodilator. Although the absence of particles in the perfusate was not confirmed (it had been filtered at 5 mm), they hypothesized that the effect was caused by the substantial content of organic compounds determined to be present. 16.4.2.5 Depression of Systemic Immune Responses There are indications from animal studies that diesel exhaust, like other combustion emissions, may suppress systemic immune function at high levels of exposure. The strongest evidence suggests suppression of T lymphocyte-mediated immunity. It is not known if environmentally relevant exposures cause this effect or which component of exhaust is responsible. The effect is not unique to diesel exhaust; it has also been demonstrated with inhaled tobacco smoke (Kalra et al., 2000) and wood smoke (Burchiel et al., 2005), all of which preferentially impact T lymphocyte function. Yang et al. (2003) instilled NIST 1650 DPM intratracheally into female B6C3F1 mice three times every 2 weeks for a 2-week or 4-week period at doses ranging from 50 to 15,000 mg/kg body weight, and examined effects in splenic lymphocytes. In mice injected with sheep red blood cells (SRBC), the number of anti-SRBC antibody-forming cells was decreased by 2 weeks of exposure in a dose-related manner. The effect was less after 4 weeks of exposure, and only significant at the highest exposure levels. The numbers of CD4þ and CD8þ T lymphocytes were reduced in the spleen by exposure, but the numbers of B lymphocytes were not reduced. The proliferative response of T lymphocytes to mitogenic stimulus was also reduced. Overall, the lowest dose causing a significant effect was 1000 mg/kg. Burchiel et al. (2004) exposed A/J mice 6 h/day, 7 days/week for 6 months to exhaust from 5.9-L 2000 Cummins engines operated on U.S. certification fuel on the heavy-duty certification transient cycle at dilutions producing DPM concentrations of 30, 100, 300, and 1000 mg/m3, followed by evaluation of effects on splenic T and B lymphocyte responses to mitogens. They found the proliferative response of T lymphocytes to be reduced similarly at all exposure levels. The response of B lymphocytes was actually increased at the lowest exposure level but not affected at other levels. The finding of suppressed T lymphocyte mitogenic response at 30 mg/m3 suggests that the potential for environmental exposures to combustion emissions to alter systemic immune function is worthy of further exploration.

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16.4.2.6 Altered Development and Growth There are indications, from both epidemiology and animal studies, that exposure to diesel exhaust might affect organ development and growth. Epidemiology provides indirect evidence that exposure to components of diesel exhaust has the potential to alter organ development, birth weight, and growth, by linking those effects to ambient pollutants that are contributed in part by diesels. The effects of exhaust have been examined directly in animal studies, but only one study has included exposure levels within the environmental “hotspot” range. Epidemiology Gauderman et al. (2000) studied 3035 children in 12 communities in southern California in the fourth, seventh, or tenth grade who had at least two spirometric evaluations over 4 years in the 1990s. In the fourth grade cohort, deficits in lung function growth were related to PM and NO2. The estimated growth rate deficit for children in the most polluted compared to the least polluted city was a reduction of 3.4% in FEV1. They obtained similar results in a follow-up study of a second cohort of 1678 children who were enrolled in the fourth grade in 1996 (Gauderman et al., 2002). There are several birth cohorts of children under followup for evaluation of the effects of prenatal exposure to urban pollutants. In New York City, inner-city mothers were enrolled during pregnancy and the children underwent prenatal assessment for exposure to PAH (Perera et al., 2006). Mothers wore a personal monitor to collect vapors and PM2.5, and umbilical cord blood was collected. In a cohort of 183 children, prenatal PAH exposure was significantly associated with a lower mental development index at age 3 and cognitive developmental delay (Perera et al., 2006). Results in 303 children from the same study indicated a combined effect of prenatal PAH exposure and postnatal exposures to environmental tobacco smoke (ETS) on cough and wheeze at age 12 months and respiratory symptoms and asthma at 24 months (Miller et al., 2004). In offspring of mothers exposed to PAHs near the World Trade Center as indicated by cord blood adducts, in combination with in utero exposure to ETS, there was a 3% reduction in head circumference and a reduction of birth weight of 8% (Perera et al., 2005). In Krakow, Poland, DNA–PAH adducts were measured in cord blood as a marker of PAH exposure, and in 333 children adduct levels were significantly related to cough, sore throat, and wheeze during the first year of life (Jedrychowski et al., 2005). In 1397 children in the Czech Republic, prenatal exposure to PM2.5 and PAHs were related to altered lymphocyte subtypes at birth in the children (HertzPicciotto et al., 2005). Ambient PM has been linked to low birth weight in other studies. In Nova Scotia, regional levels of PM10, and SO2 were linked to home address and found to be associated with lower birth weight in 74,284 births during 1988–2000 (Dugandzic et al., 2006). Other associations have been observed between PM and preterm birth (Hansen et al., 2006) and increased postnatal mortality in Southern California (Woodruff et al., 2006). None of these studies has specifically examined diesel exhaust exposure, but diesels undoubtedly contributed to the exposures. Studies in Animals Studies in animals of the effects of inhaled diesel exhaust on organ development and growth have produced mixed results. Although results to date suggest the plausibility of developmental effects, few conclusions can be drawn from the scanty research done to date. Mauderly et al. (1987b) exposed F344 rats 6 h/day 5 days/week to exhaust from 1980 5.7-L Oldsmobile engines operated on U.S. certification fuel on an urban transient duty cycle, diluted to a DPM concentration of 3500 mg/m3 (NOx ¼ 5 ppm). One group was exposed beginning with conception until 6 months of age, and a parallel group was exposed

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as adults between 6 and 12 months of age. Measurements after exposure included bronchoalveolar lavage indicators of inflammation and cytotoxicity, immune responses SRBC in pulmonary lymph node cells, respiratory function, clearance of inhaled tracer particles, lung tissue collagen, lung histopathology, and lung burdens of DPM at the end of exposure and after a 6-month recovery period. The percentage of females bearing litters was higher in the exposed group than in controls, and the number of live pups per litter was similar between the groups. Most effects of exposure were identical in the two groups. Rats exposed as adults had increased lung weight, retarded clearance of tracer particles, and considerable intra-alveolar aggregation of retained DPM. Lung weight was not increased, and tracer particle clearance was not slowed in rats exposed during development, and retained DPM was more scattered in the lung. Moreover, retained DPM was cleared more rapidly from the lung during recovery in the younger group. These results did not suggest that exposure to a high concentration of exhaust during breeding and gestation affected fertility or that exposure during lung development had greater impact than exposure during adulthood; indeed, differences tended in the opposite direction. Researchers at the Tokyo Metropolitan Institute of Public Health have conducted studies of the effects of in utero exposures to high concentrations of diesel exhaust on development of nonrespiratory organs of rats. Initial studies revealed that exposures could alter bone mass balance in growing rats (Watanabe and Nakamura, 1996; Watanabe, 1998). Attention then turned to reproductive organs. Watanabe and Oonuki (1999) exposed F344 rats 6 h/day 5 days/week for 3 months beginning at birth to whole exhaust from a 0.3-L diesel engine operated at constant speed and at a dilution containing 5630 mg DPM/m3 and 12.2 ppm NOx. A parallel group was exposed to filtered exhaust. Exposure to both whole and filtered exhaust increased serum testosterone and estradiol and decreased follicle-stimulating hormone and spermatogenesis. The only difference in responses to whole and filtered exhaust was a greater depression of lutenizing hormone by whole exhaust. Watanabe and Kurita (2001) then exposed pregnant F344 rats 6 h/day between days 7 and 20 of gestation using the same whole and filtered exhaust generation system and concentrations, and examined effects in maternal hormone levels and organ development in the near-term fetuses. Exposure to both whole and filtered exhaust increased the anogenital difference in both male and female fetuses (an indicator of “masculinization”) and retarded development of the testis, ovary, and thymus. Maternal testosterone and progesterone levels were increased by whole exhaust and decreased by filtered exhaust. Watanabe (2005) then examined effects of in utero exposures on the numbers of sperm and Sertoli cells in mature rats, using the same exhaust generation system. Pregnant F344 rats were exposed 6 h/day to whole exhaust at DPM concentrations of 1710 or 170 mg/m3 or to filtered exhaust at the same dilutions, and the testis and epididymis of the male offspring were examined at 96 days of age. Testis and epididymis weights were unaffected by exposure, but the sperm, spermatid, and Sertoli cell counts were reduced similarly by both whole and filtered exhaust at both exposure levels. This is the lowest concentration reported to cause significant effects to date and suggests that repeated exposures during pregnancy to diesel exhaust at high occupational or environmental “hot spot” concentrations might affect spermatogenesis in offspring of reproductive age. Although not a study of development, Tsukue et al. (2001) also found effects of high exposures on male reproductive organs of rats. Exposure of adult (13-month-old) F344 rats for 8 months to whole exhaust at dilutions containing 300, 1000, or 3000 mg DPM/m3 did not alter testicular weight, but even the lowest concentration reduced prostate and coagulating gland weights. The highest level increased prostate, coagulating gland, and seminal vesicle

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weights. Changes in serum hormone levels were also found. These findings support the hypothesis that occupational levels of exposure might affect male reproductive function.

16.5 CURRENT ISSUES It could be speculated in view of 60 years of continuous research that studies of the health impacts of diesel emissions could be brought to a close. A review 5 years ago (Mauderly, 2001) pointed to several issues remaining in question, and these have only been partially resolved. Moreover, recent results have raised health issues receiving little previous attention, and evolving emission standards, fuel, engine, and after-treatment technologies, and air pollutant causality issues raise additional questions. It appears that the era of diesel exhaust research is not yet completed. Four issues requiring continued study are raised by the information presented in preceding sections. The results of the new epidemiological studies of miners and truck drivers need to be completed, analyzed, and reported in detail. The incorporation of better estimates of exposure than possible in earlier studies may provide a greater confidence in estimates of unit risks for lung cancer and other health outcomes, especially if the risks derived from the two populations are similar. The cardiovascular effects of diesel exhaust require further exploration. The range of potential effects has not been explored, and better information is needed on the levels of exposure required to cause the effects. The effects of exposure on organ development and reproductive function warrant further scrutiny; again to explore the range of outcomes and whether exposure-response relationships extend down into plausible occupational and environmental exposure ranges. Finally, the relative effects of emissions from different combustion sources are still an open issue. Diesel exhaust has received attention second only to cigarette smoke, but the scanty comparisons to date suggest that diesels present few, if any, hazards that are not also presented by exhaust from gasoline or compressed natural gas engines, wood smoke, and other combustion sources. In addition to the preceding continuing concerns, the following three issues are among those begging further study. 16.5.1

Causal Components

The components of the complex mixture that cause the diverse health effects hypothesized or demonstrated to be caused by diesel exhaust remain an important issue. Determining whether or not the causal components are eliminated is of increasing importance as fuel, engine, and after-treatment technologies evolve to meet emission standards for DPM mass and NOx. For example, it has often been assumed (and sometimes explicitly stated without demonstration) that DPM caused the effects of whole exhaust, yet a growing body of knowledge indicates that this is often not true. The contributions of some subclasses of emissions are largely untested, such as the SVOCs or nonelemental carbon-based “nanoparticle” condensates. Parsing the effects among the different components is difficult and requires direct comparisons for confirmation. Demonstrating that a component can cause an effect at some dose is a useful starting point, but comparisons to whole exhaust or other components are necessary to determine whether that component predominates the effect of whole exhaust. Current knowledge of causal components is largely limited to information from studies comparing the effects of whole and filtered exhaust.

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The earliest studies of filtered versus unfiltered diesel exhaust focused on carcinogenicity in animals chronically exposed at extreme concentrations. Both Heinrich et al. (1986) and Brightwell et al. (1989) found that removing DPM eliminated the lung tumor response of rats and markedly reduced noncancer inflammation, fibrosis, and epithelial alterations as well. These findings are consistent with the understanding that developed later that the rat lung carcinogenicity was a nonspecific response to overloading with retained particles. More recent studies have addressed noncancer effects in humans and animals and have clearly shown that the nonparticle components also present health hazards. Rudell et al. (1999) found that adding a ceramic particle trap only reduced lung inflammatory responses of humans to single exposures by approximately 20%, although most of the DPM mass was removed. In contrast, Campen et al. (2005) found that filtering exhaust reduced the lung inflammatory response in mice exposed for 3 days. They found, however, that the nonparticle components accounted for all of the electrocardiographic abnormalities in the mice. Multiple Japanese studies of rodents exposed to whole or filtered exhaust, described in preceding sections, have shown that all, or the majority, of effects on amplification of allergic responses from in utero or postnatal exposures (Maejima et al., 2001; Watanabe and Ohsawa, 2002) and organ development (Watanabe and Oonuki, 1999; Watanabe and Kurita, 2001; Watanabe, 2005) were attributable to nonparticulate emissions. There have been few other studies including comparisons that point to causal components. Seagrave et al. (2001) measured inflammatory responses in rat lungs after instillation of combined PM and SVOC or the two fractions separately, collected in the heavy-duty bore of a traffic tunnel. They found that the SVOC contributed more to the response than the PM, even though there was less SVOC mass. Although both fractions caused inflammation, the SVOC fraction was four-fold more potent per unit mass than the PM. Using a similar animal response model and a multivariate statistical approach McDonald et al. (2004d) determined that hopanes and steranes, markers of engine oil, were the chemical species that covaried most closely with the inflammatory and cytotoxic effects of instilled PM and SVOC collected directly from emissions of normal- and high-emitting diesel and gasoline vehicles. These species existed primarily in the PM phase. Overall, the results to date demonstrate the fallacy of assuming, without confirmation, that any one physical-chemical fraction of diesel exhaust is responsible for the effects of whole exhaust. Results to date also show that different fractions are likely responsible for different effects. There are questions about the hazards presented by the smallest portion of DPM, falling into the loosely defined size range broadly termed “nanoparticle.” Although PM of this size has always been present in diesel exhaust and other combustion emissions, this fraction has received increased attention due to the recent upsurge of “nanotechnologies” in the manufacturing, personal product, and pharmaceutical fields, along with accompanying concerns for potential health hazards (Borm et al., 2006). Although combustion-derived nanoparticles have been discussed as a subcomponent of the more general issue (Donaldson et al., 2005), most health researchers have focused only on small (singlet) units of the ECbased soot agglomerates. It is not widely appreciated by the health community that many exhaust nanoparticles are condensates of SVOCs that have no EC core. The behavior of this material in the lung is unknown. Indeed, to the extent that the material disperses on the surfactant-rich liquid surface layer of the lung, it may not persist in particulate form, and any hazard would likely be driven by bulk chemistry in which the delivered mass would be the critical dose determinant. The delivered mass, in contrast to particle number, would be very small. There are yet no published studies of the toxicity of this class of DPM or evaluating the relative contributions of this class compared to other DPM. The issue is important because

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emissions of this material could persist if heavy organic vapors are not eliminated by new emission control strategies. Indeed, as EC is removed, this class could become a predominant form of DPM (although at very low mass emission rates). 16.5.2

Health Implications of Aged Exhaust

With very few exceptions, studies of the health effects of diesel exhaust have addressed fresh emissions. This is especially true for laboratory studies, and is also largely true for epidemiological studies of occupational exposures. While this case is important for on-road and near-road (or near-engine) exposures, people are exposed in the environment to exhaust and its atmospheric reaction products of various ages up to several days. Exhaust components are transformed by condensation of semivolatile fractions onto particles, oxidation reactions, chemical reactions with other air contaminants in the presence of temperature changes and sunlight, and other processes (reviewed by Zielinska, 2005). These changes are at least partially understood from studies involving atmospheric reaction (“smog”) chambers, but it is undoubtedly true that the full range of transformations is not known. There have been few attempts to link smog chamber studies to health assays, although the mutagenicity of DPM extracts has been shown to be altered by aging in such systems (Claxton and Barnes, 1981). In simpler reaction systems, the exposure of DPM to ozone has been shown to increase the lung inflammatory response of rats dosed by instillation (Madden et al., 2000) and production of inflammatory cytokines in cultured human airway epithelial cells (Kafoury and Kelley, 2005). The few studies of animals exposed to aged exhaust, however, have not suggested an increase in toxicity. Pepelko and Peirano (1983) exposed strain A mice 20 h/day, 7 days/week for 8 weeks to “raw” or irradiated and aged exhaust at 6000 mg DPM/m3, followed by a 9-month recovery period and evaluation of lung tumor incidence. Both the percentage of mice with tumors and the number of tumors per mouse were lower in those exposed to irradiated exhaust. Haussmann et al. (1998) exposed Wistar rats 6 h/day, 7 days/week for 13 weeks to fresh diesel exhaust or exhaust aged for 30 min and evaluated indices of inflammation and cytotoxicity in the lung. Both the numbers of inflammatory cells and the concentration of lactate dehydrogenase in lung lavage fluid were more than 50% lower in rats exposed to the aged exhaust. Questions about the effect of aging and atmospheric transformations on the health hazards from diesel exhaust blend quickly into the larger issue of the physical-chemical species responsible for the health effects of air pollution in general. However, a careful assessment of the health impacts of diesel exhaust must consider that effects of fresh and aged exhaust might differ. 16.5.3

Health Impacts of Reduction of Emissions and New Technologies

Nearly all of our information on the health impacts of diesel exhaust pertains to fuel, engine, and after-treatment technologies that are already or are rapidly becoming outdated. The information from laboratory studies also results from a wide spectrum of on-road and offroad engines spanning 25 model years and an assortment of fuels (many not specified). While it is true that the longevity of diesel engines ensures that older technology systems will remain in use for some time, it is also true that emissions from new vehicles and equipment bear scant resemblance to emissions from even 10 years ago. There is little doubt that compression ignition will continue to comprise a portion of internal combustion engines for the foreseeable future. These engines will undoubtedly continue to be called “diesels,” but

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the marked changes in technology and resulting emissions warrant a distinction between emissions from systems marketed today for on-road use (and rapidly including off-road systems as well) and those of the past. There is yet no straightforward way to demarcate old engine versus new engine emissions, although Hesterberg et al. (2005) coined the term “new technology diesel” (NTD) to describe a combination of advanced engine design, ultra-low sulfur fuel, specialized lubricants, and catalyzed particle traps. While it is logical, and undoubtedly correct, that “less is better” when it comes to health hazards from engine emissions, the health benefits of tightened emission standards and the residual hazards that may remain are worthy of continued evaluation. At this time, there is only one published study directly comparing the health effects of inhalation exposure to conventional diesel and NTD exhaust. McDonald et al. (2004b) exposed C57BL/6 mice 6 h/day for 7 days to the same dilution of exhaust from a 0.4-L singlecylinder engine operated in steady state at full load under two conditions. Conventional conditions were simulated by fueling the engine with contemporary average on-road fuel (certification fuel at 371 ppm sulfur), and NTD conditions were simulated by using ultra-low sulfur fuel (ECD1, British Petroleum, 14 ppm sulfur) and adding an appropriately sized catalyzed ceramic particle trap. The DPM mass, CO, and VOC emissions were markedly reduced under the NTD condition, but NOx emissions were unchanged (no NOx reduction after-treatment was used). Evaluations of health response included lung histopathology, clearance of Respiratory Syncytial Virus, and inflammatory cytokines and indicators of oxidative stress in lung homogenate. All indicators were significantly changed by conventional exposure, but no significant differences from control were induced by NTD exposure. These results suggest that reduction of DPM, VOC, and CO emissions by improved fuel and after-treatment will likely have substantial health benefit, even if the benefits, which will only be implemented gradually, will be difficult or impossible to confirm by epidemiology. It is also useful to know that these benefits were achieved by fuel and after-treatment strategies that can be retrofitted to existing conventional systems. A related issue is whether or not alternative fuels or advanced after-treatment strategies might introduce unintended consequences in the form of health hazards. Numerous new liquid fuels are entering the diesel fuel portfolio; some are already in use, and others are on the way. All pundits forecast at least partial displacement of conventional petroleum diesel stocks by these fuels. These include petroleum-water emulsion blends, vegetable and animal oil-derived biodiesel, grain and biomass-derived alcohols, North American tar sand petroleum, gas-to-liquid and coal-to-liquid fuels, dimethyl ether, and various blends of these with conventional petroleum fuel. Few of these fuels have been tested for health hazards, and there have been even fewer comparisons to conventional petroleum-based fuel. In compliance with EPA requirements for certifying new fuels and fuel additives, 90-day inhalation studies of rats exposed to multiple levels of exhaust from engines burning soybean oil-derived biodiesel (100% soy methyl ester) (Finch et al., 2002), petroleum diesel–water emulsion (Reed et al., 2005), and the same emulsion containing methanol (Reed et al., 2006a) were conducted with numerous health endpoints. Few significant effects were found for any of the fuels even at the highest exposure level (they were subsequently approved for commercial use), but two issues remain. First, none of the studies included comparison to conventional diesel fuel. Second, the regulatory test protocol does not address several health outcomes that have resulted from high exposures to conventional diesel exhaust, such as cardiovascular effects, effects on allergic responses or systemic immune function, or effects on organ development. Exhaust from engines burning the other new fuels has received almost no scrutiny. The scanty emissions data available and the fact that the new fuels will be used in

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combination with new engine and after-treatment technologies suggest that new health issues are unlikely; however, some attention to comparative toxicity screening is certainly warranted.

ACKNOWLEDGMENTS Mauderly’s effort was supported by the National Environmental Respiratory Center, a joint government-industry program funded by the EPA Office of Research and Development (CR831455-01-0), the U.S. Department of Energy (DOE) FreedomCar and Vehicle Technology Program (DE-FC04-96AL76406), the DOE Office of Fossil Energy (DE-FC2605NT42304), and several corporations in the automotive, electrical utility, engine, and petroleum sectors. This chapter has not been reviewed by sponsors, and is not intended to reflect the views of any sponsor. Garshick’s effort was supported by the Office of Research and Development, Department of Veterans Affairs, National Institutes of Health National Cancer Institute (R01 CA90792), and National Institute of Environmental Health Science (R21 ES013726).

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17 DIOXINS AND DIOXIN-LIKE CHEMICALS Michael A. Gallo and Morton Lippmann

17.1 INTRODUCTION Dioxins, a subset of the polyhalogenated aromatic hydrocarbons (PHAHs), are ubiquitous in the environment at low concentrations. At much higher concentrations, they are known to produce a variety of adverse health effects in laboratory animals, and interspecies modeling suggests, to some public health professionals, that they pose exceptionally high health risks to people exposed to background levels. The extents of the public health risks have been the subject of a series of risk assessments conducted by the United States Environmental Protection Agency (EPA) since the mid-1980s (NRC, 2006), but EPA has not yet issued a formal risk assessment. In 1995, and again in 2001, EPA’s Science Advisory Board (SAB) issued reports that evaluated draft EPA risk assessments for dioxin and dioxin-like compounds but did not endorse them as being scientifically valid (SAB, 1995, 2001). The most recent EPA draft reassessment was reviewed by a National Research Council (NRC) committee, which also failed to fully endorse its methodology and its quantitative estimates of public health risks (NRC, 2006). This chapter includes an update of the chapter authored by Michael J. DeVito and Michael A. Gallo for the second edition of this book on the science of dioxin and dioxin-like chemicals (DLCs), along with a summary of the findings and conclusions of the NRC Committee and a review of some recent papers on the associations of exposures to dioxin and DLCs in human populations and health effects. As noted above, dioxins and DLCs are a subset of the PHAHs. The most well studied and the most toxic of the class is 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). These chemicals are present in a variety of environmental media as well as low-level contaminants of the food supply. The dioxin-like PHAHs consist in part of the polyhalogenated dibenzo-p-dioxin, dibenzofurans, and biphenyls (PCBs and PBBs). These chemicals are sparingly soluble in water and are highly lipophilic. They are persistent in both environmental and biological samples, with half-lives in humans ranging from 1 to over 20 years (Flesch-Janys et al.,


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1995). The dioxin-like chemicals induce similar toxicities in experimental animals. These toxicities are initiated, in part, by the binding and activation of these chemicals to an intracellular protein called the aryl hydrocarbon receptor (AhR) The health effects of dioxin and related chemicals in humans and wildlife are hotly debated with the intensity of the debate based as much on the uncertainties in risk assessment as it is for political, social, and economic reasons. Despite being one of the most studied class of chemicals, risk assessments for dioxin by various governmental health and regulatory agencies throughout the world have resulted in tolerable daily intakes or virtually safe doses that range over almost three orders of magnitude. This large discrepancy is due to the use of either threshold or linear models in cancer risk estimates. The political and social issues revolve, in part, around the widespread background exposure to the “most toxic man-made chemical” as well as to the use of dioxin-contaminated herbicides during the Vietnam War. In addition, there have been a number of industrial accidents that resulted in human and wildlife exposure to these chemicals, and attempts to resolve the exposures and effects of these accidents have been widely criticized from all sides. The debate on the health effects of dioxin and related chemicals is not likely to be resolved in the near future.

17.2 SOURCES One of the problems in estimating potential health effects of dioxin is the coexposure to numerous dioxin-like chemicals originating from different sources. There are 75 different polychlorinated dibenzo-p-dioxin (CDDs), 135 dibenzofurans (CDFs), and 209 biphenyls (PCBs), depending upon the number and position of the chlorine substitutions. Fortunately, only a subset of these chemicals produces dioxin-like toxicities in experimental animals. The 2,3,7,8-substituted CDDs and CDFs are considered dioxin-like. Only 7 of the 75 CDDs and 10 of the 135 CDFs are dioxin-like (Van den Berg et al., 1998). Of the 209 PCB congeners, only 12 have dioxin-like activity (Safe 1994; Van den Berg et al., 1998). In addition, brominated analogues have been found in the environment as well as in human tissues. The data available on the brominated analogues demonstrates that these compounds induce dioxin-like biochemical effects, toxicities, and similar structure–activity relationships with their chlorinated analogues (Safe 1990). Because dioxins consist of a broad class of chemicals with varying potencies, sources, and exposures, health risk assessments have used the toxic equivalency factor (TEF) methodology. This methodology is a relative potency scheme that compares the toxicity of all dioxins to the most potent, that is, TCDD). Congeners are assigned relative potency values called TEFs (Van den Berg et al., 1998). The concentration of a congener in a mixture is multiplied by its TEF, and this product is the TCDD or toxic equivalents (TEQ) for that congener. The TEQs for all chemicals in the mixture are summed to produce a TEQ for the entire mixture. It is assumed that the mixture will behave as if it contained the TEQ concentration of TCDD alone. At present, only 29 chemicals are considered in the TEF scheme, and they consist of the 2,3,7,8-substituted CDDs/CDFs and 12 out of 209 PCBs (Van den Berg et al., 1998). TEF values for the brominated analogues have not been adopted by regulatory agencies. The development of this methodology is described in greater detail later in this chapter. The CDDs and CDFs are unwanted products of several industrial processes. Combustion systems are a primary source for the production of CDDs and CDFs. Included in this category are waste incinerators, such as municipal solid waste, medical waste, sewage sludge, and hazardous waste incinerators. The burning of fuel, such as coal, wood, and petroleum, also

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produce CDDs and CDFs. Other high-temperature sources, such as cement kilns, also produce significant amounts of dioxins. Iron ore sintering, steel production, and scrap metal recovery operations also produce and release CDDs/CDFs into the air. In some parts of the United States, open burning of household trash is a common practice, and it results in the formation of CDDs/CDFs. The contribution of open burning of household trash to total emissions of dioxin into the environment cannot be quantified with any certainty (USEPA, 1997). Several chemical manufacturing processes result in the formation of CDDs/CDFs as by-products, such as the production of herbicides, such as Agent Orange or 2,4,5-trichlorophenoxy acetic acid, and chlorinated phenols such as pentachlorophenol. These processes have either been altered to eliminate dioxin production or the products have been discontinued. The manufacture of chlorine-bleached wood pulp results in the production of trace quantities of CDDs/CDFs, of which the octa- and heptachlorinated congeners predominate. It should be noted that many pulp and paper mills have reengineered their processes and have decreased CDD/CDF production and emissions in these facilities by approximately 90% (Cleverly et al., 1998). The production of ethylene dichloride or vinyl chloride produces CDDs/CDFs; however, the data are insufficient to quantify emission estimates for these processes. Several natural processes can result in the production of CDDs/CDFs. The predominant congeners are hepta- and octachlorodibenzo-p-dioxin produced in forest fires. Under certain environmental conditions, such as composting, microorganisms may produce CDDs/CDFs from chlorinated phenolic compounds. Ball clay deposits in western MS, KY, and TN were found to contain CDDs and CDFs. The CDDs/CDFs in these clay deposits were approximately 90% by weight 2,3,7,8-tetrachlorodibenzo-p-dioxin and 1,2,3,7,8-pentachlorodibenzo-p-dioxin. While this congener pattern is similar to that found in contaminated herbicides, the origin of these chemicals in the clay has not been determined and natural occurrence is but one possibility. Another source of CDDs/CDFs is redistribution and circulation of reservoirs of previously released CDDs/CDFs (USEPA, 1994). Contaminated sediments, soils, and pentachlorophenol-treated wood are also considered reservoirs. Retrainment of these reservoirs may contribute significantly to overall exposure, but its exact contribution is uncertain. While there are natural sources for CDDs/CDFs production, several lines of evidence indicate that recent emissions and exposures are due to anthropogenic sources. Analysis of sediment cores in the United States and Europe shows consistent patterns of CDD/CDF concentration changes over time. CDD/CDF concentrations in sediment tend to increase starting around the 1920s–1930s, peaking between the 1960s and the 1970s, and then beginning to decrease in the later 1970s through the 1980s (Czuczwa et al., 1985; Cleverly et al., 1996). The increasing trends are consistent with the increase in general industrial activity in the 1930s, and the current decreasing trends are attributable to the promulgation of environmental regulations since the 1970s (Czuczwa et al., 1985). CDD/CDF concentrations are higher in human tissues from industrialized countries compared to those from underdeveloped nations (Schecter et al., 1994b, 1994c). There are also data comparing present human tissue concentrations with those tissues taken from preserved 140–400-year-old human remains that show almost the complete absence of CDD/CDFs compared to tissues of modern man (Schecter et al., 1994b, 1994c). Finally, no known large natural sources of CDD/CDFs have been identified. Estimates of all emission sources suggest that forest fires are a minor source compared to anthropogenic sources (Cleverly et al., 1998).

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The EPA developed an inventory of CDDs/CDFs sources for the years 1987 and 1995 (Cleverly et al., 1998). The year 1987 was chosen because prior to this year few potential sources in the United States had been characterized, and 1995 was chosen as the latest year for which significant data were available. This analysis indicates that the dominant source of dioxin releases to air in the United States is combustion. The estimate of releases in 1987 was 12 kg TEQ (range 5–30 kg) and from 1995 was 3 kg TEQ (range 1–8 kg). While there was uncertainty in these estimates, it was clear that emissions significantly decreased between 1987 and 1995. The reductions in emissions were due primarily to reductions in air emissions from municipal and medical waste incinerators (Cleverly et al., 1998). An updated EPA inventory for releases into the environment of PCDDs, PCDFs, and TCDD-like PCBs for the United States for 2000 was published in 2005 and was included in the NRC (2006) review. There was a reduction of 89% from a 1987 best estimate. Some dioxin-like chemicals were synthesized and sold commercially, such as the polychlorinated and polybrominated biphenyls, PCBs and PBBs, respectively. The PCBs were used as heat transfer fluids; flame-retardants; paint additives; and dielectric fluids for capacitors, transformers, and in several other industrial processes (Devoogt and Brinkman, 1989). The PBBs were used predominantly as flame-retardants in the early 1970s. Since 1929, approximately 1.5 million metric tons of PCBs were produced and sold. Numerous commercial mixtures of PCBs were manufactured and sold worldwide, including Arochlor, Clophen, Fenclor, Kanechlor, Phenoclor, and Pyralene among others. These mixtures were sold as blends based on their chlorine content. For example, Aroclor 1242, the most widely produced Aroclor, contained 42% chlorine while Aroclor 1254 contained 54% chlorine. In the United States, the sale of PCBs was banned in 1977. Since 1977, the disposal of PCBs has been strictly regulated under the Toxic Substances Control Act (TSCA). If disposals of PCBs following TSCA guidelines are strictly followed, environmental release should be minimal. The majority of current PCB releases would appear to be from rerelease of these compounds from reservoir sources. Other current sources of PCBs are most likely due to leaks and spills of still in-service PCBs, from transformers, for example, or from illegal disposal of PCBs. Despite the ban on the production of PCBs in the United States and Western Europe in the 1970s, PCBs were not banned in the former Soviet Union until the 1990s. 17.2.1

Environmental Fate and Transport

Much of the information on the fate and transport of CDDs/CDFs is based on the data for TCDD. Because CDDs/CDFs share many physical and chemical properties, the fate and transport of these chemicals should be qualitatively similar. CDDs/CDFs enter the terrestrial food chain via atmospheric deposition (Fries and Paustenbach, 1990). Recent airborne sources of dioxin are dominated by the combustion of wastes and fuels. The CDDs/CDFs emitted from combustion sources are predominately bound to particulate matter (PM), although there are some in the vapor phase. Once airborne, the CDDs/CDFs deposit on plants, soil, or water. Plant contamination by dioxins is mainly due to the deposition of contaminated PM on leaves. While there is evidence that CDDs/CDFs in soil can be absorbed directly by the roots of plants, it is highly unlikely that this results in significant concentrations of the dioxins in the above-ground plant (Insensee and Jones, 1971; Jensen et al., 1983). CDDs/CDFs also accumulate in soil through pesticide application and leakage from waste sites, although these pathways are more important for localized areas with specific problems. Once deposited on the plants or soil, the dioxins either enter the food chain or degrade. The deposition of dioxins on plants and soil enter the food chain through direct ingestion of the

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plants or the incidental ingestion of soil by animals, resulting in the bioaccumulation of these chemicals in livestock (Fries and Paustenbach, 1990). Concentrations of CDDs/CDFs are higher in cow’s milk collected from the vicinity of a municipal waste incinerator than in commercial cow’s milk (Rappe et al., 1987). Dioxins can be degraded in the environment. However, this appears to be a relatively slow process. Photolysis is the main degradation pathway for CDDs/CDFs in the environment and requires ultraviolet (UV) light and an organic hydrogen donor (Crosby et al., 1971). Because of the requirement for UV light, dioxins that are on soil surfaces have shorter half-lives than those that are deeper in the soil. For example, the half-life of TCDD at the soil surface is estimated at 1–3 years, while it is 10–12 years in the subsurface soil (di Domenico et al., 1982; Kimbrough et al., 1984). Dioxins are sparingly soluble in water, and the concentrations of these chemicals in water are extremely low. Despite their low solubility, aquatic environments have significant amounts of dioxin as a result of their adsorption to sediments. Sediments of surface waters are thought to be the ultimate sink (or environmental reservoir) of CDDs/CDFs (Hutzinger et al., 1985), and the persistence of these compounds in water bodies results in bioaccumulation in aquatic organisms (Insensee and Jones, 1975). While deposition of PM contaminated with dioxin appears to be the major source of water contamination, local events such as industrial effluents and herbicide runoff may also contribute to the dioxin burdens. The PCBs and PBBs have similar physical and chemical properties as the CDDs/CDFs, which result in qualitatively similar environmental fates and transports (Safe, 1994). Atmospheric concentrations of PCBs in urban centers (1–10 ng/m3) were approximately an order of magnitude higher than in nonurban areas (0.1–0.5 ng/m3) (Atlas et al., 1986). Numerous factors including local sources, source emission strengths, and meteorological conditions influence ambient concentrations of PCBs. Surface waterways appear to be a major reservoir for PCBs (Tanabe and Tatsukawa, 1986), and the National Academy of Sciences (NAS, 1979) estimated that 50–80% of the PCBs in the environment are contained within the waterways of the North Atlantic. Similar to the CDDs/CDFs, PCBs are predominantly found in the sediments of these waterways. PCBs are also found in soil to varying degrees from 0.01 to over 2000 ppm (Tatsukawa, 1976). Photolysis and photoxidation appear to be the major pathways for destruction of PCBs in the environment. All PCBs photodechlorinate, and the photolysis rate increases as the chlorine content increases (Tiedje et al., 1993). In addition to photolysis, PCBs are also sensitive to biological degradation. There are over two dozen strains of aerobic bacteria and fungi that are capable of degrading most PCB congeners with five or fewer chlorines (Tiedje et al., 1993). Many of these organisms are of the genus Pseudomonas or Alcaligenes, and they are widely distributed in the environment. The higher chlorinated PCBs are more resistant to biodegradation than are the lower chlorinated congeners. In addition, PCBs substituted in two or more of the ortho positions are resistant to biodegradation. 17.2.2

Human Exposure

Exposure to dioxin can occur through occupational exposure, accidental exposures, or environmental exposures to the general population. The predominant hypothesis for environmental exposure to the general population focuses on the air to plant to animal hypothesis. This hypothesis focuses on the deposition of PM onto plants and soil. The deposited dioxins then enter the food chain by ingestion of the contaminated substrates by either livestock or aquatic life where they bioaccumulate. Eventually, the livestock or aquatic lives are consumed by humans.

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Estimates of daily human intake in the United States in the 1990s were approximately 3–6 pg TCDD equivalents/kg/d (Pinsky and Lorber, 1998; Schecter et al., 1994a), and these values are consistent with those reported from Western Europe (Schecter et al., 1994b, 1994c). Approximately half of the TCDD equivalents arose from PCBs, with only 10% attributable to TCDD itself (Schecter et al., 1994a, 1994b, 1994c). These daily intakes resulted in serum concentrations of approximately 40–60 parts per trillion (ppt) TEQ, and body burdens of approximately 8–13 ng TEQ/kg (DeVito et al., 1995). In general, dioxin body burdens increase with age, and this is thought to be the reflection of higher exposure of both during the 1940s through the 1960s compared to more recent exposures (Pinsky and Lorber, 1998), as well as decreased metabolism of dioxin with age (Flesch-Janys et al., 1995). A major route of elimination of dioxin in female humans, experimental animals, and mammalian wildlife is through breast milk. Breast milk has a high concentration of fat, and the dioxin distribute to fatty tissues. In women, nursing for at least seven months can decrease maternal serum concentrations of dioxin by up to 50% (Schecter et al., 1996). Exposures to nursing infants are much higher than maternal exposure and is estimated at 30–100 pg TEQ for the first year (Schecter et al., 1994a, 1994b, 1994c, 1996). These daily intakes are higher than all tolerable daily intakes defined by regulatory agencies throughout the world. In nursing infants, body burdens of dioxin at 12 months are two to four times higher than maternal burdens (Schecter et al., 1996). The potential health risks associated with these background exposures are uncertain; however, it should be noted that the benefits of breastfeeding are well documented and far outweigh the potential risks from background dioxin exposures (Rogan and Gladen 1993). There have been numerous incidents throughout the world where small populations have potentially been highly exposed to dioxin through industrial accidents, occupational exposures, wartime use of herbicides, or environmental pollution. One of the most wellcharacterized exposures occurred in Seveso, Italy in 1976. Approximately 1 kg of TCDD was released following an explosion at a trichlorophenol manufacturing plant. In 1976, there were no validated methodologies available to determine serum concentrations of TCDD. Despite this fact, a team of physicians, lead by Dr. Paolo Mocarelli, collected and stored blood from over 30,000 patients in the area, both exposed and unexposed (Bertazzi and di Domenico, 1995). Initial chemical analysis of the serum from several children from the most highly exposed area found serum concentrations of dioxin up to 50,000 ppt (Bertazzi and di Domenico, 1995). Based on the initial exposure estimates, the Seveso region was divided into three areas A, B, and R. Region Awas thought to be more highly exposed than region B, and region R was the unexposed area. More recent characterization of the exposures in Seveso indicates that the average serum concentrations in regions A are lower than the initial studies indicated (Bertazzi and di Domenico, 1995). Several industrial cohorts have been examined for dioxin exposure. Most of these workers are either farm workers spraying phenoxy herbicides or workers manufacturing herbicides or trichlorophenol (Fingerhut et al., 1991; Flesch-Janys et al., 1995). Workers in these studies have had serum concentrations of dioxin ranging 100–5000 ppt TEQ on a lipid-adjusted basis. TCDD is the predominant congener in the occupational exposures, in contrast to the low-level background exposure of the general population, where TCDD contributes approximately 10% of the total TEQ. Spraying of herbicides contaminated with TCDD during the Vietnam War exposed military personnel from both sides of the conflict, as well as Vietnamese civilians. Exposures of members of the U.S. Armed Forces are the best characterized from these groups. Despite

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initial fears of high exposure during the war, it appears that, with the exception of those directly involved in the formulation and the spraying, most ground troops had limited exposure to these chemicals. The spraying occurred predominantly in the south of Vietnam. Civilians from South Vietnam had approximately 10 times the serum concentrations of TCDD compared to those in North Vietnam (Schecter et al., 1994b, 1994c). Concentrations of dioxin were similar between residents of South Vietnam, the United States, and Western Europe (Schecter et al., 1994b, 1994c). Accidental exposures to dioxins and dioxin-like compounds have occurred following consumption of contaminated rice oils. In 1968, a mass poisoning, called Yusho or oil disease, occurred in Fukuoka and Nagasaki prefectures in Japan due to ingestion of rice oil contaminated with Kanechlor-400, a commercial PCB mixture produced in Japan. It was subsequently discovered that the Kanechlor-400 mixture was contaminated with CDFs as well as polychlorinated tetraphenyls (Masuda, 1995). A similar incident occurred in Yucheng, Taiwan, in 1979 (Hsu et al., 1995). These incidents are known as Yusho and Yu-cheng, respectively. The concentrations of dioxin in the Yusho population shortly after the accident were approximately 20 times higher than controls, while PCB concentrations were 20–50 times the background (Masuda, 1995). Similar exposures were observed in the Yu-cheng incidents as well. While these populations were intensively studied, it has been difficult to determine which of the effects seen were due to the dioxin, the non-dioxin-like PCBs, or to their coexposures. 17.2.2.1 Pharmacokinetics The dioxin-like toxicity of these chemicals is due to the parent compound, and for the CDDs and the CDFs, metabolism is a detoxification step. Dioxins have relatively long half-lives in biological systems. In rats, the half-life of dioxin ranges from approximately 1 day for TCDF and 3,30 ,4,40 -tetrachlorobiphenyl to greater than 6 months for octachlorodibenzo-p-dioxin (OCDD) (Van den Berg et al., 1994). TCDD has a half-life ranging from 9 days in mice (Birnbaum, 1986; Gasiewicz and Rucci, 1984; Gasiewicz et al., 1983) to approximately 1 year in rhesus monkeys (Bowman et al., 1989). In humans the half-life of the CDDs and CDF range from months to over 20 years. Estimates of the half-life of TCDD range from 5.7 to 11.3 years, with an average of about 8 years (Flesch-Janys et al., 1998; Pirkle et al., 1984). The difference in the half-life between mice, rats, and humans is approximately 80–400-fold. Such large differences in half-life among species have significant impact on the dose metric used in comparing species sensitivity to these chemicals. Many of the earlier risk assessments were based on comparisons of daily dose or administered dose and did not adequately correct for the difference in the kinetics of these chemicals among species. More recent comparisons have used steady-state body burdens as the dose metric, which provided a more accurate comparison, although it has limitations depending upon the end point of comparison (DeVito et al., 1995; Wang et al., 1997). The term half-life must be used cautiously with dioxin for several reasons. First, in humans there is a relationship between percent body fat and half-life. The greater the percent body fat the longer the half-life (Flesch-Janys et al., 1995, 1998). In addition, there is nonlinear kinetics of these chemicals that precludes the use of a single half-life. The disposition of dioxin is dose dependent due to hepatic sequestration (DeVito et al., 1997; Andersen et al., 1993). The liver contains CYP1A2, which is inducible by dioxin. CYP1A2 binds TCDD and sequesters dioxin in the liver. In CYP1A2 knockout mice, neither TCDD nor 4-PeCDF is sequestered in hepatic tissue (Diliberto et al., 1997). In the wild-type mice, the induction of CYP1A2 results in the sequestration of these chemicals in the liver.

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The dose-dependent hepatic sequestration has been demonstrated in rats, mice, hamsters, guinea pigs, and humans (Van den Berg et al., 1994; Carrier et al., 1995). While TCDD is sequestered in the liver, other dioxins are sequestered to a greater degree. Because of these nonlinearities in the kinetics of dioxin, using a single half-life value has limitations, is best used as an indication of the persistence of these chemicals, and may not be suitable for quantitative determination of exposures. Intensive efforts were made to develop physiologically based pharmacokinetics models for dioxin. Most of these models focused on TCDD (Andersen et al., 1993, 1997; Kedderis et al., 1993); however, a few have examined other congeners such as 2,3,7,8-tetrabromodibenzo-p-dioxin, dibenzofurans, and PCBs (De Jongh et al., 1993; Kedderis et al., 1993). These models not only describe the pharmacokinetics of dioxin but are also the basis for response models examining gene transcription and hepatocarcinogensis (Kohn et al., 1996; Conolly and Andersen 1997; Portier et al., 1996). These models have potential use in risk assessment but require further validation. Congener specific pharmacokinetics information for PCBs in humans is also available. Those chemicals that have been examined are persistent, with half-lives in humans ranging from months to years (GE data). However, there are significant differences between the metabolism and disposition of PCBs compared to CDDs and CDFs. While metabolism for the CDDs and CDFs is a detoxifying pathway, metabolism of some of the dioxin-like PCBs produces bioactive metabolites. For example, hydroxylation of PCBs 77, 105, and 118 results in metabolites that bind to transthyretin, one of the thyroxine-binding proteins in serum. It has been hypothesized that the binding to transthyretin displaces thyroxine and increases its elimination (Brouwer et al., 1998). Other hydroxylated PCBs bind to utero globulin and accumulate in tissues expressing this protein, particularly the lung. Similar to the CDDs and CDFs, PCBs 77, 126, and 169 are sequestered in hepatic tissue. In contrast, the mono-ortho dioxin-like PCBs do not accumulate in the liver (Devitor et al., 1998). There is also evidence of pharmacokinetics interactions between CDDs, CDFs, and PCBs (van der Plas et al., 1998; Van den Berg et al., 1994), and extrapolations of animal data on single congeners to human exposures must be viewed with caution.

17.3 TOXICOLOGICAL EFFECTS AND MECHANISMS OF ACTION In experimental animals, dioxins induce numerous toxicities including immunotoxicity, reproductive and developmental toxicities, and carcinogenicity (DeVito and Birnbaum, 1995; Safe, 1990; Pohjanvirta and Tuomisto, 1994). The lethal effects of dioxin are unique in that death occurs weeks after the initial exposure and is preceded by a wasting syndrome. In acute exposure studies, the time to death appears independent of the dose, and increasing the dose does not decrease the time to death. Studies by Rozman and coworkers determined that the lethal effects of dioxin are time dependent (Viluksela et al., 1997). Animals initially die from the wasting syndrome within weeks following the initial exposures. However, animals that survive the initial wasting syndrome may eventually die from other causes, such as anemia, months after the initial exposure. The time to death, while independent of dose, can be altered by hypophysectomy. Lethal effects of TCDD can be observed within the first 24 h in hypophysectomized mice (DeVito et al., 1992a). While the dose response and time course for TCDD-induced wasting syndrome and lethality have been well characterized, the exact cause of death is uncertain. The wasting syndrome is thought to be due to alterations in the body weight set point. While the wasting syndrome can be severe, it does not appear to be

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a direct cause of death. Pair-fed controls with the same weight loss as the TCDD-treated animals did not exhibit mortality (Seefeld et al., 1984). 17.3.1

Carcinogenicity

The carcinogenicity of TCDD has been examined in rats, mice, hamsters, and Japanese medaka and is positive in all four species (Huff et al., 1994). Several studies in rats and mice indicated that TCDD is carcinogenic when administered for the lifetime of an animal. Tumors were observed in both sexes at multiple sites including liver, thyroid, lung, and several other tissues (Huff et al., 1994). In hamsters, the species most resistant to the lethal effects of TCDD, epidermal tumors were observed after dosing once a month for 6 months by either oral gavage or subcutaneous injections (Rao et al., 1988). Few of the related dioxins have been examined for carcinogenicity in 2-year bioassays. A mixture of HxCDDs induced liver tumors in female rats and thyroid tumors in the males. The Aroclors 1016, 1242, 1254, and 1260 mixtures have been tested for carcinogenicity, and all mixtures induce liver tumors while 1242, 1254, and 1260 also increase the incidence of thyroid tumors (Mayes et al., 1998). TCDD is negative in several short-term mutagenicity assays. In addition, using methods that can detect one DNA adduct in 1011 nucleotides, no TCDD-derived adducts have been detected (Turtletaub et al., 1990). Carcinogenesis is a multistage process that requires discrete steps involving genetic alterations of cells that clonally expand and progress into tumors. In this multistage process, TCDD clearly acts as a tumor promoter, although it may act on multiple stages of this process. TCDD is one of the most potent tumor promoters (Pitot et al., 1980; Poland et al., 1982; Lucier et al., 1991). All of the dioxin-like chemicals tested in hepatic tumor promotion models have tested positive, including 1,2,3,7,8-pentaCDD, 2,3,4,7,8-pentaCDF, and PCBs 126, 169, 118, 105, and 156 (Hemming et al., 1993, 1995; Buchmann et al., 1994; Maronpot et al., 1993; Haag-Gronlund et al., 1997, 1998; Schrenk et al., 1994). Much of the research on the carcinogenic effects of dioxin has examined the liver tumors in rats. In rats, the development of hepatic tumors occurs only in the female (Kociba et al., 1978; NTP, 1982). Tumor promotion studies in ovariectomized rats indicate that TCDD does not promote liver tumors in rats without a functioning ovary (Lucier et al., 1991). However, in these studies, the ovariectomized rats developed lung tumors, while the intact rats only developed hepatic tumors (Lucier et al., 1991). It has been hypothesized that TCDD increases the metabolism of estrogens and results in the production of catechol estrogen metabolites (Lucier et al., 1991). These metabolites are thought to redox cycle, resulting in the production of oxygen free radicals that then produce DNA damage (Tritscher et al., 1996). While the role of estrogen is critical in the TCDD-induced liver tumor in female rats, it should be noted that hepatic tumors were not observed in hamsters, and in mice males were more responsive to the hepatic carcinogenic effects of TCDD than were females (NTP, 1982). In rats and mice, TCDD is an extremely potent carcinogen. The LOAEL for hepatic adenomas in rats is 10 ng/kg/d, while the NOEAL is 1 ng/kg/d (Kociba et al., 1978; NTP, 1982). These doses seem rather high compared to the human intake of 1–6 pg TEQ/kg/d. To make an appropriate comparison, the large difference in half-life between species must be taken into account. One method is to express dose as steady-state body burdens (ng/kg). Daily intake of a chemical will eventually result in a steady-state condition in which the amount ingested equals the amount eliminated. With persistent chemicals such as dioxin,

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small concentrations ingested daily can result in large accumulation of the chemical in the body. In rats, 1 ng/kg/d results in a body burden of approximately 30 ng TCDD/kg body weight. In mice the LOAEL for hepatocarcinogenesis is 71 ng/kg/d (NTP, 1982), with a resulting steady-state body burden of 944 ng/kg (DeVito et al., 1995). 17.3.2

Developmental Toxicity

TCDD and several other congeners induce cleft palate and hydronephrosis in mice. Cleft palate can be induced in rats and hamsters, but only at doses that result in significant fetal mortality (Olson and McGarrigle, 1991). Studies in cultured developing palates indicate that mouse palates are approximately 100–1000 times more sensitive to the effects of TCDD than are human and rat palates (Abbott and Birnbaum, 1991; Abbott et al., 1994). In addition, human and rat palates are equally sensitive to the effects of TCDD, suggesting that humans are unlikely to develop cleft palate from dioxin. Other more subtle effects of dioxin have also been observed in developing animals. The developing reproductive system of rats and hamsters is extremely sensitive to the actions of TCDD and other AhR agonists. Prenatal exposure to TCDD decreases epididymal sperm counts in mice and epididymal and ejaculated sperm counts in rats and hamsters (Theobald and Peterson, 1997; Mably et al., 1992a, 1992b; Gray et al., 1995). Female rats and hamsters exposed to TCDD in utero and lactationally develop malformations of the phallus, clitoris, and incomplete opening of the vaginal orifice (Gray et al., 1995, 1997). These developmental alterations of the reproductive systems can occur at doses as low as 0.05 mg/kg when administered on gestational day 15 (Gray et al., 1995). In a multigenerational study, doses as low as 1 ng/kg/d decrease fertility in the F1 and F2 generation (Murray et al., 1979). The 1 ng/ kg/d dose in the multigenerational study resulted in steady-state body burdens of approximately 30 ng/kg (DeVito et al., 1995). Dioxin-like chemicals are also developmental neurotoxicants. Prenatal exposure to TCDD produces a permanent low-frequency auditory deficient in rats (Goldey et al., 1996). Similar effects were observed in animals prenatally treated with either Aroclor 1254 or PCB 126 (Goldey et al., 1996; Crofton and Rice, 1999). The development of the auditory system depends upon thyroid hormones (Goldey et al., 1996). TCDD and other dioxin decrease circulating thyroid hormones by inducing uridine diphosphate glucuronsyl transferase (UDPGT) (Henry and Gasiewicz, 1987). In rats, pre- and postnatal exposure to TCDD and other dioxins decreases circulating thyroid hormone concentrations during the period of cochlear development, particularly the regions of the cochlea responsible for lowfrequency hearing (Goldey et al., 1996; Goldey and Crofton, 1998; Crofton and Rice, 1999). The auditory deficits in rats produced by TCDD is a relatively high-dose phenomena requiring at least 1 mg/kg on gestational day 18, which is approximately 20 times higher than the doses needed to alter the developing reproductive tract. The developmental neurotoxicity of TCDD is also expressed as a permanent change in regulated body temperature in both rats and hamsters (Gordon et al., 1995, 1996). Several laboratories have demonstrated behavioral changes in animals exposed prenatally to dioxinlike chemicals. Schantz and Bowman (1989) examined the neurological developmental effects of TCDD in rhesus monkeys. Rhesus monkeys were exposed prenatally and lactationally to TCDD with dams receiving as little either 5 or 25 ppt of TCDD in the diet. The offspring had alterations in object learning at the low dose, and at the high dose few of the dams were able to maintain pregnancy and only one offspring from seven dams survived (Schantz and Bowman, 1989). The rhesus monkeys fed a diet of 5 ppt TCDD averaged

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151 pg/kg/d (Bowman et al., 1989), and this diet resulted in a body burden of approximately 42 ng/kg (DeVito et al., 1995). 17.3.3

Immunotoxicity

TCDD and related chemicals are immunotoxicants in several species (Harper et al., 1993). The immunotoxicity of dioxin is difficult to characterize, and there does not appear to be a unique dioxin-like immune response. The responses affected by TCDD depend on the species studied and the model used to examine the immune response. For example, the suppression of the plaque forming cell response to sheep red blood cells in mice is one of the most consistent findings, with an ED50 of approximately 0.7 mg TCDD/kg (Smialowicz et al., 1994). Yet in rats, TCDD enhances the response to sheep red blood cells (Smialowicz et al., 1994). In contrast, exposures to the same doses of TCDD that enhance the response to sheep red blood cells suppress host resistance to Trichinella spiralis (Luebke et al., 1994, 1995). In influenza models, doses of TCDD as low as 10 ng/kg increase mortality in mice following exposure to influenza virus (Burleson et al., 1996). Other low-dose effects on the immune system by TCDD are altered lymphocyte subsets in marmoset monkeys exposed to 0.3 pg/kg/week for 24 weeks (Neubert et al., 1992). This dose results in body burdens of approximately 10 ng/kg (DeVito et al., 1995). The immunotoxicity is thought to be associated with changes in proliferation and differentiation in a variety of cell types in the immune system (Kerkvliet, 1995).

17.4 MECHANISMS OF ACTION 17.4.1

Role of the Ah Receptor in the Biological Effects of Dioxin

Binding to and activating the AhR is the initial step in the biological and toxicological effects of dioxin. Most, if not all, of the effects are mediated by this protein (Birnbaum 1994). The unliganded AhR is found in either the cytosol or nucleus as a multimeric complex that includes two molecules of a 90 kDa heat shock protein and several other smaller molecular weight proteins (Whitlock et al., 1996). Upon ligand binding, the AhR dissociates from this complex and binds to the aryl hydrocarbon nuclear translocator (ARNT). The transformed AhR–ARNT complex then binds to specific dioxin-responsive enhancer (DRE) sequences located in the promotor region of the CYP1A gene and several other TCDD responsive genes. The binding of the activated AhR complex to the DREs alters chromatin structure and enhances the association of other components of the transcriptional machinery resulting in the initiation of transcription. This transformation of the AhR into a nuclear binding form can be attenuated by protein kinase C inhibitors, and phosphorylation events appear to be important in regulating the activity of these proteins (Whitlock et al., 1996). While the exact role of the AhR in normal biochemical and physiological processes is uncertain, there are several lines of evidence that suggest that it plays an important function in developmental and homeostatic functions. AhR, a member of the beta-helix-loop-helixPer-ARNT-Sim (bHLH-PAS) superfamily, is a ligand-activated transcription factor. The AhR is a highly conserved protein present in all mammalian species examined and has been found in all vertebrate classes examined, including modern representatives of early vertebrates, such as cartilaginous and jawless fish (Hahn, 1998). In addition, a possible

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AhR homolog has been identified in Caenorhabditis elegans (Powell-Coffman et al., 1998). The bHLH-PAS superfamily consists of at least 32 proteins found in diverse organisms from Drosophila, C. elegans, and humans. These proteins are transcription factors and appear to require dimerization, either homo-or heterodimers, for functional effects. These proteins regulate circadian rhythms (per and clock), steroid receptor signaling (SRC-1, TIF2, RAC3), or are involved in sensing oxygen tension (HIF-1, EPAS-1/HLF) (Hahn, 1998). It has been proposed that understanding the function of the bHLH-PAS family of proteins and the phylogenetic evolution of the AhR may lead to an understanding of the role of this protein in normal processes. Other lines of evidence indicate that the AhR has important physiological functions based on its spatial and temporal expression in developing embryos. The AhR is expressed in tissue- and cell-specific manner during development (Abbott et al., 1994, 1995, in press). It is highly expressed in the neural epithelium, which forms the neural crest during development (Abbott et al., 1995). AhR knockout mice have been produced using a targeted disruption of the Ahr locus (Fernandez-Salguero et al., 1997). While the results from the two laboratories are not identical, it is apparent that the AhR is not essential to life, since the knockout mice survive and reproduce. However, the Ahr/ mice develop numerous lesions with age (Fernandez-Salguero et al., 1997). Mortality begins to increase at about an age of 20 weeks, and by 13 months 46% of the mice either died or were ill. Cardiovascular alterations consisting of cardiomyopathy with hypertrophy and focal fibrosis, hepatic vascular hypertrophy and mild fibrosis, gastric hyperplasia, T-cell deficiency in the spleen, and dermal lesions were apparent in these mice, and the incidence and severity increased with age (Fernandez-Salguero et al., 1997). While male and female AhR-/-mice are fertile, the females have difficulty maintaining conceptuses during pregnancy, surviving pregnancy and lactation, and rearing pups to weaning. In one study, only 46% of 39 pregnant AhR-/-females successfully raised pups to weaning (Abbott et al., in press). Other groups have developed an ARNT knockout mouse. In contrast to the Ahr/ mice, the lack of its dimerization partner, ARNT, results in fetal mortality. Unlike the AhR, ARNT has several other dimerization partners including HIF-1 alpha (Hahn, 1998). It has been suggested that some of the toxicity of AhR ligands such as TCDD may be due to sequestration of ARNT by the AhR, decreasing ARNTs availability for other members of the PAS family (Hahn, 1998). The importance of the AhR in the toxicity of dioxin stems from several lines of evidence. First, structure–activity relationships indicate a correlation between receptor binding affinity and in vivo toxicity in mice (Safe, 1990). Second, there is genetic evidence in mice that is consistent with the AhR mediating the toxic effects of dioxin. The C57BL mice are the most sensitive to the toxic effects of dioxin while the DBA/2J is the most resistant. The binding affinity of the AhR is approximately 10–20 times lower in the C57BL compared to the DBA, and the C57BL mice are approximately 10–20 times more sensitive than the DBA mice (Poland and Glover, 1980). In fact, the sensitivity to dioxin in mice segregates nicely with several different alleles and the binding affinity of dioxin to these gene products (Poland and Glover, 1990). Finally, the most powerful evidence is that the AhR/ mice are resistant to the toxicities of TCDD (Mimura et al., 1997; Fernandez-Seguaro et al., 1996). While the initial step in the toxicity of dioxin is the binding to the AhR, this binding is clearly not the sole determinant in the toxicity of these chemicals. In the mouse, the different AhR alleles have different binding affinities to dioxin, and these differences can explain the differences in sensitivity to dioxin between these strains (Poland and Glover, 1980). The biochemical and biophysical properties of the AhR are similar between species, with only


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slight differences in molecular mass and subunit stability (Gasiewicz and Rucci, 1984). Yet for some end points, such as lethality, there are significant differences in the ED50 values between species (Pohjanvirta and Tuomisto, 1994). While the ED50 values for some end points, such as fetal mortality and enzyme induction, are similar between the species (Olson and McGarrigle, 1991). Overall, these data indicate that the steps after AhR binding and activation are critical in understanding the species differences in response to dioxin. 17.4.2

Dioxins as Growth Dysregulators and Endocrine Disruptors

While dioxins appear to alter numerous systems, there are several underlying commonalities of the toxicity of these chemicals. Most of the toxicities are associated with alterations in proliferation and/or differentiation, such as cancer, immunotoxicity, and chloracne. Hyperplastic responses are observed in gastric mucosa and bile ducts in monkeys (McConnell and Moore, 1979), urinary bladder in guinea pigs, and hepatic and dermal hyperplasia occurs in several species. Hypoplasia occurs in lymphoid tissues in all mammalian species examined and in the gonads of mice, rats, rabbits, and mink. Squamous metaplasia is induced by TCDD in ceruminous and sebaceous glands in monkeys (Moore et al., 1979). Dysplastic responses of the nails and teeth have been observed in humans and nonhuman primates following prenatal exposure to dioxin (Moore et al., 1979; Masuda, 1995). Dioxins are potent growth dysregulators. Dioxins alter the normal homeostatic processes by disrupting cell signaling pathways through multiple mechanisms. Cells maintain homeostasis through a complex processes involving the release of paracrine or autocrine hormones or growth factors. These hormones and growth factors interact with specific receptors that can be localized to the cell membrane, cytosol, or nucleus. Activation of these receptors initiates a cascade of events that affect cell functioning. Dioxins alter cell signaling and homeostasis by altering hormones and growth factors, and their receptors and by altering the signaling of the activated receptors. These effects of dioxin are often tissue and developmental stage specific (DeVito and Birnbaum, 1995). TCDD alters many of the signaling pathways involved in the endocrine system as well and can be considered an endocrine disruptor. Dioxin alters serum concentrations of several pituitary hormones (Moore et al., 1989) and steroid hormones including androgens (Moore et al., 1985), glucocorticoids (Lin et al., 1991), and thyroid hormones (Henry and Gasiewicz, 1987; Kohn et al., 1996). In cell culture, TCDD increases estradiol metabolism by almost 100-fold (Spink et al., 1990). Estrogen receptors are decreased in several tissues including liver and uteri in mice and rats administered dioxin (DeVito et al., 1992a, 1992b; Safe et al., 1991). Dioxins have frequently been described as antiestrogens because they decrease estrogen and estrogen receptor concentrations. However, this characterization should be viewed cautiously, since dioxins also increase the incidence and severity of endometriosis in monkeys (Reier et al., 1993) and increase endometriotic lesions in mice and rats (Cummings et al., 1996; Johnson et al., 1997). Endometriosis is dependent upon estrogens for growth and is often treated with drugs that downregulate ovarian function. Once again, this highlights the tissue-specific effects of these chemicals. 17.4.3

Human Health Effects Associated with Dioxin Exposure

The epidemiological data examining the potential carcinogenic effects of dioxin have focused mostly on several industrial cohorts of pesticide manufacturers (Kogevinas

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et al., 1997), American veterans of Vietnam, and the Seveso, Italy, cohorts. It has proved difficult to obtain a consensus on the interpretation of these studies. IARC evaluated the evidence for carcinogenicity of TCDD and other dioxins (McGregor et al., 1998). The IARC working group focused on the four cohorts with the highest exposure (Fingerhut et al., 1991; Hooiveld et al., 1998; Ott and Zober 1996; Flesch-Janys et al., 1995; Zober et al., 1994). In these studies, the overall risk of developing cancer was significantly greater (approximately 1.4-fold) in the exposed populations (McGregor et al., 1998). However, few site-specific cancer risks were consistent across the cohorts, but evidence of increased risk of lung cancer was significant in all the four studies (Fingerhut et al., 1991; Hooiveld et al., 1998; Ott and Zober 1996; Flesch-Janys et al., 1995; Zober et al., 1994). One of the problems with interpretation of these studies is that an increase in risk for cancers at many sites was observed for only a few cases, as is the case for smoking exposure to ionizing radiation in atomic bomb survivors (McGregor et al., 1998). In contrast to TCDD, both smoking and ionizing radiation also increase site-specific tumors. Because TCDD is the only chemical that increased risk for all cancers without consistently increasing the risk for specific tumors suggests that the epidemiological data should be viewed with caution. The IARC workgroup considered the evidence for carcinogenicity as limited. Another problem with the interpretation of the cancer data for the most highly exposed occupational cohorts is the lack of a consistent exposure–response relationship. This can be seen in Fig. 1 from a paper by Starr (2002) that shows the data from the studies of workers in Hamburg (Flesch-Janys et al., 1998), the NIOSH study (Fingerhut et al., 1991), and for the BASF cohort (Ott and Zober, 1996). It can be seen that the SMRs above unity in almost all of the exposed groups show little, if any, association with TCDD body burden over a range covering three orders of magnitude. It is likely that the excess cancers in these groups of chemical plant production workers were attributable to their occupational exposures to many other toxic chemicals that were produced in those plants or to their tobacco smoking. Bodner et al. (2003) extended the analysis of the 2187 Dow Chemical Company workers in the NIOSH cohort by 9–30 years of follow-up. This cohort included workers with high exposures, as evidenced by 11% of them having had clinically confirmed chloracne. There was no excess for all cancers, and an SMR for lung cancer of 0.8. Also, there was no increased cancer risk with increasing exposure. In a recent study of all workers in the NIOSH cohort at eight U.S. plants with sufficient data to estimate their exposure to TCDD, Cheng et al. (2006) sought to determine whether the association between TCDD and cancer was due to, or modified by, other occupational exposures. They used both untransformed and ln-transformed TCDD ppt-years, lagged 15 years in their analyses, as well as in their analyses of smoking-related cancers and other cancers. They reported that their estimated incremental lifetime mortality risk at age 75 from all cancers was 6–10 times lower than previous estimates for this cohort that did not consider the age and concentration dependence of TCDD elimination. In view of the relatively weak epidemiological data on cancer in humans, the IARC workgroup also considered mechanistic information to support their overall evaluation that TCDD is carcinogenic in humans (McGregor et al., 1998). In 2005, after completion of its dioxin reassessment, EPA revised its cancer guidelines and charged the NRC Committee to use the new guidelines in addressing the scientific evidence for classifying TCDD as a human carcinogen. In its report (NRC, 2006), there was a split among the committee members on whether the evidence met all of the criteria necessary for definitive classification of TCDD as “carcinogenic to humans,” although they could all agree that it should be considered “likely to be carcinogenic to humans.”


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Considering the somewhat equivocal evidence of TCDD carcinogenicity among highly exposed industrial workers, and the currently much lower and still dropping levels of exposure in the general population, the cancer risks to the general public from contemporary environmental levels of dioxin and DLCs are extremely small. In fact, there is suggestive evidence that very low levels of dioxin exposure may be associated with reduced risks of cancer based on the observed rates in the NIOSH occupational cohorts (Kayajanian, 2002). The results are not inconsistent with an examination of the association between benign prostatic hyperplasia (BPH) and serum dioxin concentrations that was made for participants in the National Health and Nutritional Examination Survey 1999–2000 (NHANES) by Gupta et al. (2006). After age adjustment, men without BPH had 21% (CI: 5–39%) higher TEQs compared to men with BPH. The population was relatively small, but it was a national probability sample. Immune effects of dioxin are observed in all species examined (Kerkvliet, 1995). However, human health has not been as consistently altered. Some of the best evidence comes from developmental studies in humans, which have suggested that there are developmental effects following exposure to CDDs CDFs, and PCBs (Weisglas-Kuperus et al., 1995). In addition, neurodevelopmental effects have been reported for a number of cohorts (Rogan and Gladen, 1991, 1992; Jacobson and Jacobson, 1997; Jacobson et al., 1990). Recent follow-up studies on people heavily exposed to TCDD in Seveso reported that plasma IgG levels decreased with increasing plasma TCDD, while there was no association between TCDD and IgM, IgA, C3, or C4 levels (Baccarelli et al., 2002). Also, for women who were heavily exposed to TCDD in Seveso, there was no clear evidence of alterations in ovarian function. The Dutch studies of a cohort of infants has demonstrated relationships between high CDDs, CDFs, and PCBs exposure and decreases in thyroid hormones and delayed neurodevelopment (Koopman-Esseboom et al., 1994; Weisglas-Kuperus et al., 1995). While these studies indicate that exposure to this class of chemicals may be associated with alterations in the developing immune and neurological systems, these effects were subclinical, and extrapolation to the population at large is uncertain. However, it should be stressed that in the Dutch studies, this cohort is of women and infants with no known high exposures to PHAHs other than background exposures. The associations found in this cohort, while subclinical, warrant further research and concern. 17.4.4

Toxic Equivalency Factors

Estimating the risk associated with dioxin exposure can be problematic. These chemicals are present as trace contaminants of a complex mixture containing numerous dioxin-like chemicals. Initially, risk assessments focused on TCDD alone, and all other dioxins were excluded. However, experimental evidence clearly indicates that these other dioxin-like chemicals cannot be ignored. If risk assessments are to include other dioxins, then they must either assume that all dioxin-like chemicals are as potent as TCDD or they must use some sort of relative potency scaling methodology. The experimental data clearly demonstrate that assuming all dioxin are equally potent to TCDD would greatly overestimate the potential risk. Hence, a relative potency scheme described as toxic equivalency factors was proposed to estimate the potential health risk following exposure to dioxin-like chemicals (Eadon et al., 1986). The toxicological basis for the TEF methodology is the shared mechanism of action of these chemicals, namely, activation of the AhR (Safe, 1990; Birnbaum and DeVito, 1995). The relative potency of a dioxin-like chemical is compared to TCDD, one of the most

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potent AhR ligands studied. The relative potency for a chemical from a particular study is described as a REP. A TEF is assigned by an expert group, using all the available REP values for a particular chemical. The data available for assigning a TEF value for a chemical varies widely depending upon the congener. For example, there are over 25 studies that examined the relative potency of PCB 126 compared to TCDD with end points ranging from in vitro studies examining binding affinity and biochemical alterations to in vivo subchronic studies examining toxicological and biochemical effects (Van den Berg et al., 1998). For other chemicals, such as PCB 80, there are only in vitro studies suggesting a dioxin-like effect (Van den Berg et al., 1998). In determining the TEF value, the varying types of data are weighted according to relevancy. For example, data on binding affinity are weighted less than data from acute in vivo studies, which are weighted less than data from subchronic studies examining toxicological effects (Van den Berg et al., 1998). TEFs are considered order-of-magnitude estimates for several reasons, including the quality and quantity of the data available, the variability of the data, and the limited information on the species extrapolation of the relative potency of these chemicals (Van den Berg et al., 1998). The relative potency of a chemical compared to TCDD is a function of its binding affinity to the AhR and its comparable pharmacokinetics properties, including absorption, disposition, metabolism, and elimination (DeVito et al., 1997;). One example of the importance of pharmacokinetics in the relative potency of these chemicals is OCDD. Originally, OCDD was assigned a TEF value of 0 because it demonstrated no toxicological effects in acute studies in rodents and binding data were unavailable due to the compound’s insolubility in aqueous solutions. However, in subchronic studies in rats, OCDD demonstrated significant dioxin-like effects and was assigned a TEF value of 0.001. In short-term studies, very little OCDD was absorbed by the animals. However, because of its persistence, in longer-term studies OCDD will accumulate and eventually attain concentrations in the animal sufficient to produce toxicological effects (Birnbaum and Couture, 1988; Couture et al., 1988). The TEF methodology assumes that the biochemical and toxicological effects of mixtures of dioxin can be predicted using a dose addition methodology. Several investigators have examined the TEF methodology, using either simple mixtures containing 2, 3, or 4 congeners, or more complex mixtures from environmental samples or laboratory prepared mixtures. The TEF methodology adequately predicted the immunotoxicity (Silkworth et al., 1989), lethality (Eadon et al., 1986), or biochemical changes (Van den Berg et al., 1989) of complex environmental mixtures. Schrenck and coworkers used the TEF methodology to predict the tumor promotion potency of a laboratory prepared mixture of 49 different chlorinated dioxins. Rozman and coworkers examined the effects of a quarternary mixture of chlorinated dioxin and found that the TEF methodology predicted both biochemical and toxicological responses of the mixture (Viluksela et al., 1998a, 1998b). While there are a number of studies demonstrating additivity and supporting the TEF methodology, there are still some unanswered questions. Recently, the TEF methodology was expanded to include the coplanar PCBs and the mono-ortho-substituted PCBs. These chemicals, while demonstrating dioxin-like toxicities, have significant non-dioxin-like effects. For example, TCDD and other chlorinated dibenzo-p-dioxins decrease thyroxine concentrations by only 40–50% in rats. In contrast, PCBs 77 and 118 can produce a 90% decrease in thyroxine (Crofton et al., 1998). In a complex mixture consisting of two chlorinated dibenzo-p-dioxins, four chlorinated dibenzofurans, and six chlorinated biphenyls, the TEF methodology predicted the immunotoxicity and enzyme inducing effects of


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the mixture but underpredicted the thyroxine decreases and porphyrin accumulation in rats administered the mixture compared to rats receiving TCDD. One of the problems with the TEF methodology is the confusion as to what this method actually does. The TEF methodology only examines dioxin-like chemicals and assumes that there are no interactions between the dioxin and the nondioxin present in the mixture. In essence, the method allows one to estimate the potential health risk of a complex mixture by assuming that only TCDD is present. This method neither makes attempt to predict interactions of dioxin and other classes of chemicals nor the potential for toxicity unrelated to the AhR-binding cascade. Hence, the method ignores potential nonadditive interactions, both antagonistic and greater than additive interactions. While there is uncertainty in the use of the TEF methodology, one question to ask is whether using this method decreases or increases our uncertainty in the overall risk assessment. As described above, either excluding all other dioxin-like chemicals or considering other dioxin-like chemicals as potent as TCDD increases uncertainty in risk assessments. At a workshop on the use of TEFs in ecological risk assessment sponsored by the USEPA, the participants clearly agreed that the use of the TEF methodology decreases uncertainty. However, quantifying the uncertainty in the TEF methodology, and its effects on the uncertainty of the overall risk assessment, remains elusive (National Research Council 2006). With no alternative methodology, one can argue that the TEF method decreases uncertainty in the overall risk assessment. Until alternative methodologies are developed, the TEF methodology is the most appropriate method to estimate the potential health risks of exposure to dioxin-like chemicals. Significant efforts should be made to clarify the uncertainty of this methodology to support risk assessors and managers until more accurate methods are developed. 17.4.5

Risk Characterization

Dioxins induce numerous toxicities in experimental animals. The toxicity of these chemicals is mediated through an interaction with the Ah receptor. The Ah receptor is a highly conserved protein throughout evolution. While the exact role of the Ah receptor is uncertain, inappropriate activation of this receptor by dioxins produces tissue-specific alterations in differentiation, proliferation, and apoptosis. These processes are critical in numerous biological systems including development, the immune and endocrine systems, and carcinogenesis, to name a few. While there may be outliers, either resistant or sensitive, to certain dioxin toxicities, most species respond at similar exposures (i.e., within an order of magnitude for dose). The available data suggest that for biochemical and some toxic responses, humans have similar sensitivities to experimental animals. For example, the steady-state body burdens in rodents that result in cancers are between 300 and 900 ng/kg (DeVito et al., 1995). In the epidemiological studies, the high end of exposures have body burdens ranging 400–5000 ng/kg (DeVito et al., 1995). There is no firm evidence to suggest that humans are less sensitive to the potential effects of dioxin than are laboratory animals. Because of the consistency of effects observed in both humans and experimental animals, and the evolutionary conservation of the structure and function of the Ah receptor, efforts should be made to limit exposure to this class of chemicals. Since the available human data have not been considered to provide an adequate basis for quantitative risk assessments for dioxins in the past, there has been a reliance on the toxicity of dioxin in laboratory animals as a basis for understanding and estimating the risks

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associated with exposure to these chemicals. In the past, regulators have relied on either linearized models for cancer risk assessments or on safety factor methods. EPA has used a margin of exposure (MOE) method to examine potential risks from these chemicals (DeVito et al., 1995). Recent human background exposure was approximately 8–13 ng TEQ/kg. Responses in experimental animals have been observed at body burdens ranging 5–900 ng/ kg or more. In the low dose range, immunotoxicity in rodents and monkeys, reproductive effects in multigeneration studies, and endometriosis in rhesus monkeys occurred at body burdens between 5 and 64 ng/kg. If one divides the body burden in animals where effects are observed, by the average human body burden one obtains the MOE. From the above data, the MOE ranges from less than 1 to approximately 10 (an order of magnitude). The MOE does not imply that effects are occurring in the population. However, it compares the difference in present human exposure to that of animals exposed to toxic doses of a chemical. Typically, the EPA considers a MOE of 100 reasonable for noncarcinogens. The World Health Organization (WHO) has also assessed the potential human health effects of dioxin. The WHO experts suggested that the tolerable daily intake be set at 1–4 pg TEQ/kg/d. They also noted that present exposures throughout the world range 2–6 pg TEQ/ kg/d. The WHO experts emphasized the Dutch cohorts and stressed that these data suggest that background exposures are associated with effects. In addition, the working group also noted that concentrations of dioxin are decreasing in the environment over the last two decades. The risk characterizations for dioxins conducted by EPA have indicated significant public health risks from environmental exposures, especially from dietary sources. Because of the seriousness of its concerns, and those of the public, EPA has requested reviews of its risk assessments for this class of chemicals from its External Science Advisory Board (SAB, 1995, 2001) and more recently from the National Research Council (NRC, 2006). All three of these in-depth reviews by broadly constituted panels with expertise covering the various scientific and medical fields have withheld endorsements of the risk estimates contained in the successive EPA risk assessments. The basic problem has been that the risks associated with exposure to dioxin have been based on (1) studies conducted with animal models; (2) require interspecies extrapolations; (3) high-to-low dose extrapolations; (4) inclusion of multiple safety factors; and/or (5) reliance on overly conservative risk models for carcinogenesis. Thus, their usefulness for risk characterization remains uncertain, at best. As noted by NRC (2006), a major issue in the designation of carcinogenicity is that “analysts must extrapolate well below the doses observed in the studies to consider typical human exposure levels. This extrapolation involves two critical decisions: (1) selecting a point of departure (POD), which corresponds to the lowest dose associated with observable adverse effects within the range of data of a study, and (2) selecting the mathematical model used to extrapolate risk from typical human exposures that are well below the POD.” The NRC Committee concluded that EPA’s decision to rely solely on a default linear extrapolation model lacked scientific support. It recommended that EPA provide risk estimates using both nonlinear and linear methods to extrapolate below PODs and that it communicate the scientific strengths and weaknesses of both approaches so that the full range of uncertainty generated by modeling of the data is conveyed in their reassessment. With regard to noncancer effects, the NRC Committee faulted the decision of the EPAs not to derive a reference dose (RfD), its traditional metric for such effects. An RfD provides (1) estimates of the proportion of the population with intakes above the RfD; (2) detailed assessment of population groups, such as those with elevated exposures; and (3)contributions of the major food and other environmental sources for those with high intakes. Without such data, the

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18 ENDOCRINE ACTIVE CHEMICALS: BROADENING THE SCOPE* Kathryn R. Mahaffey, Shirlee W. Tan, K. Christiana Grim Jessica C. Meiller and Donald R. Bergfelt

18.1 INTRODUCTION In the early-to-mid 1990s, significant evidence that chemicals were affecting the endocrine system of wildlife drew attention from both the public and scientific communities. Because of this, intensive laboratory-based studies were designed to evaluate chemicals for potential endocrine activity across taxonomic groups (among others see U.S. EPA, 1998; Reiter et al., 1998; Cooper and Kavlock, 1997; Colborn et al., 1993). Initial research emphasized the potential of chemical agents to alter the reproductive process that primarily reflected earlier experiences in reproductive toxicology (Cooper and Kavlock, 1997). More than a decade later, attention has expanded to a broader domain of endocrine effects such as cellular and molecular effects, multigenerational effects, and many of the physiological/pathological responses such that endocrine disrupting chemicals have become elucidated. Increasingly, it is recognized that the investigation of environmental chemicals cannot rely solely on utilization of in vitro systems (e.g., isolated hormone receptors), but rather it is the use of these data with real-world exposure, epidemiological, and in vivo testing data from whole animals that depicts a more accurate picture of the impact of chemical exposures on the endocrine system. The endocrine system is an elegant regulatory network that is necessary for maintaining physiological homeostasis, and which reacts to internal (e.g., blood pressure) and external cues (e.g., daylight through the pineal gland). It is composed of organ systems that interact with one another through hormone signals, which are, in turn, regulated through highly intricate feedback mechanisms. Generally, the components of the endocrine system include neuroendocrine/endocrine regulatory feedback mechanisms and hormone production, *

The findings and conclusions in this chapter are those of the authors and not of the U.S. Environmental Protection Agency. The chapter is not being formally disseminated by U.S. EPA, so it should not be construed as representing any Agency determination or policy.


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ENDOCRINE ACTIVE CHEMICALS: BROADENING THE SCOPE

Regulatory/feedback mechanisms Neuroendocrine system

}

}

Peripheral endocrine system

A Hypothalamus Releasing factors

B Pituitary

-Interfere with regulatory mechanisms (positive and negative feedback) of circulating hormone concentrations -Interfere with potentiators of hormone regulation *Proper function of the neuroendocrine system requires the same components as the peripheral system: hormone production, transport, conversion, recognition, internalization, action, and metabolism. Therefore, endocrine interference can occur within the neuroendocrine system through similar mechanisms as those listed below.

Stimulating factors

C Hormone production

D Hormone transport

E Hormone conversion

-Interfere with signal to initiate hormone synthesis -Upregulate or downregulate hormone synthesis -Alter normal hormone structure hence altering function -Interfere with storage and release of hormones

-Interfere with production of binding/transport proteins -Change binding affinities to serum and cytoplasmic transport proteins -Alter ratio of specific binding proteins available to natural hormones

-Interfere with normal conversion of hormones to active/inactive forms -Interfere with conversion of hormones to other hormones (e.g., steroidogenesis pathway)

-Compete with membrane and nuclear receptors of hormones -Up-/downregulate production of hormone receptors -Act as a mimic or antagonist of natural hormones -Change in hormone receptor structure and function F -Activation/deactivation of postreceptor pathways Hormone recognition, internalization and Action -Interfere with postreceptor pathways (DNA transcription/translation and protein synthesis) and action of endogenous hormones -Epigenetic changes (e.g., changes in methylation patterns that influence gene expression) to target cells increasing or decreasing sensitivity of cells to natural hormones, or permanently changing genetic material passed to subsequent generations resulting in latent effects

G

Metabolism and elimination

-Interfere with hormone biotransformation, conjugation, and elimination

H Latent and population effects

FIGURE 18.1

Points of potential endocrine activity of chemical compounds.

transport, conversion, receptor recognition, and internalization, target organ action, and metabolism and elimination (Fig. 18.1). These endocrine components are vastly conserved over vertebrate taxa. Hence, research and information on hormonal activity, regulation or disruption in one vertebrate species can generally be applied to other vertebrate species.

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Concerning endocrine disruption, often the most definitive information available to make causal relationships between environmentally induced endocrine activity and physiological alteration is elucidated through controlled studies in laboratory animals and wildlife. These animal studies, coupled with the high homology between endocrine function and regulation across vertebrate species, allow greater confidence when making causal relationships between exposure to endocrine toxicants and disease in humans. While it is convenient to break down the endocrine system into distinct entities for specific functions for comprehension, such as the hypothalamic–pituitary–thyroid (HPT) axis, the hypothalamic–pituitary–gonadal (HPG) axis, or the hypothalamic–pituitary–adrenal (HPA) axis, the reality is that all components of the endocrine system are integrative, and all react and respond to one another. For instance, regulation of the reproductive system is dominated by the hypothalamic–pituitary–gonadal axis, which controls sexual maturation, gamete maturation, ovulation and pregnancy, to name a few. However, thyroid hormone is integral to regulating reproductive cyclicity in some species (Ben Saad and Maurel, 2004), hence interference in thyroid hormone action can ultimately affect the “central” reproductive axis and subsequently the reproductive capacity of an individual. In addition, several hormones have multiple and supportive functions and may be called to action simultaneously to regulate several different processes to accomplish an end result. These interrelationships within and outside of the endocrine system are important to consider when assessing the effects of endocrine active chemicals (EACs), or chemicals that interact with and influence endocrine function, either directly or indirectly on clinical disease. Human and animals are exposed to chemicals through several pathways. These exposure pathways are no different for EACs than for other chemicals. Examples of EAC exposure pathways include chemicals that travel through air such as mercury and dioxin; agrochemicals or veterinary use chemicals and hormones applied on farms and in concentrated animal feeding operations (CAFOs); and the use of pharmaceuticals and personal care products that enter our sewage treatment plants, effluents, surface waters, seawaters, groundwater, and some drinking waters (Fent et al., 2006). Much of the initial focus on EACs was on estrogen or estrogen mimicking chemicals (Segner et al., 2003), and some of the most potent of the EACs are considered to be natural and artificial steroid estrogens of anthropogenic origin, which are commonly detected in sewage effluents (Johnson et al., 2007). The biological potency of these differ from, and are influenced by, various biodegradation processes (D’Ascenzo et al., 2003). Depending on the type of sewage treatment, estrone and estradiol can be removed from as little as 5% to 14% to as much as 85% to 96% (Braga et al., 2005). These hormones, along with pharmaceuticals and personal care products, are introduced into the aquatic environment primarily by both untreated and treated sewage (Daughton and Ternes, 1999). Given the abundance of possible exposure pathways in the environments of humans and animals, it is critical that methods of detection for potential effects of these chemicals be developed and utilized. However, such a task is not straightforward. For example, the endocrine effects produced by pharmaceuticals in water typically involve species not targeted by the original investigators who developed the chemicals. Consequently, evaluation of wildlife species, ex situ or in situ has often not been part of the original assessment (Fent et al., 2006). Also problematic is that many chemicals were tested only in in vitro systems. Serious drawbacks of such systems, as well as those of applied quantitative structure–activity relationships (QSARs) have been noted (Fent et al., 2006). Likewise, the problem of mixtures is rarely addressed. Widespread environmental contamination has been documented in several other countries including France, Germany, Canada, Brazil, and the United States (Cargouet et al., 2004; Ternes et al.,

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1999; Boyd et al., 2003). Thus, the major task is to understand how EACs affect species and how such effects can be detected. EACs can interfere at several levels within an organism (Fig. 18.1), either directly (e.g., xenoestrogens interacting with estrogen receptors) or indirectly (e.g., through cytotoxicity of nonendocrine organs leading to endocrine dysfunction), potentially with the same outcome. They can also induce latent, irreversible effects that are manifested in adulthood according to the concepts of developmental origins of health and disease (DOHaD; Lau and Rogers, 2004; and in subsequent generations through transgenerational effects (Campbell and Perkins, 1988; Anway and Skinner, 2006). The distinction between the latent effects of chemical exposure on the offspring associated with DOHaD and transgenerational modes of action (MOA) is that, in the former, exposure of the pregnant parent results primarily in prenatal exposure through an in utero/in ovo route whereas, in the latter exposure of one or both parents occurs before mating or spawning such that the effects are thought to be passed along to the offspring through genomic or epigenomic routes. Such effects are difficult to study because they challenge the conventional approach to toxicology and risk assessment. They can be permanent, self-sustaining, are not necessarily dose-dependent, nor influenced by threshold levels of EACs, and they may not be expressed until years, or generations after the insult occurred, making conventional risk assessment difficult and regulation reliant on reproducible assay systems. Latent and transgenerational manifestations of disease associated with DOHaD, and the role that epigenomics plays in understanding causal relationships between exposure to endocrine toxicants and latent disease, are emerging areas of research within the field of endocrine toxicology. In this review, we (1) provide an overview of potential end points and clinical signs associated with exposure to EACs in myriad species using thyroid hormone as an example; (2) provide an example of a classic environmental contaminant, in this case inorganic lead, and its influence on a clinically important endocrine system that is rarely considered in a paradigm that focuses on reproductive biology, specifically the vitamin D/endocrine system; and (3) address the new and exciting disciplines of DOHaD and epigenetic and transgenerational effects of EACs using current examples from humans, laboratory species, and wildlife. 18.2 BIOMARKERS: TERMINOLOGY FROM VARIOUS DISCIPLINES There are myriad disciplines that are involved in understanding the effects of EACs in humans and wildlife. Due to the multidisciplinary nature of this field, there are also differences in terminology, especially as related to the term biomarker. Interestingly, in the literature associated with human epidemiology (among others, see discussion by Munoz and Gange, 1998), medicine, and health, the term biomarker (especially biomarkers of exposure, biomarkers of effect, and biomarkers of susceptibility) has arisen from terminology used by the World Health Organization (for a summary report, see WHO, 2001) and by various committees organized under the National Research Council of the National Academy of Sciences (NRC/ NAS) in the United States (among other see NAS, 2006). The term is defined as follows: Biomarker of exposure: The chemical or its metabolite or the product of an interaction between a chemical and some target molecule or cell that is measured in a compartment in an organism. Biomarker of effect: A measurable biochemical, physiologic, behavioral, or other alteration in an organism that, depending on the magnitude, can be recognized as associated with an established or possible health impairment or disease.

BIOMARKERS: TERMINOLOGY FROM VARIOUS DISCIPLINES

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Biomarker of susceptibility: An indicator of an inherent or acquired ability of an organism to respond to the challenge of exposure to a specific chemical substance. This terminology is often used in studies evaluating an association between exposures to environmental chemicals and health outcomes (typically adverse) that could include, but are in no manner limited to, endocrine effects. Also, older literature exists that utilized biomarkers of exposure (e.g., blood lead monitoring) and principally reports on assessments conducted as part of occupational and public health surveillance. Thus, the general terms for biomarkers, as these terms are used in public health, environmental epidemiology (Munoz and Gange, 1998), and environmental health refer to biological indicators (NAS, 2006) including behavior, biochemical, molecular, genetic, immunologic, or physiologic signals of events in biologic systems. These events are considered (NAS, 2006) to follow a continuum between indicators of external exposure to a chemical and resultant clinical effect (Schulte and Perera, 1993). In addition, MOA is not necessarily a critical feature of biomarkers, as used in human public health, epidemiology, and surveillance. Alternatively, there are other definitions for the term biomarkers in fields of study that focus on the ecological effects of environmental chemicals on populations of nondomesticated animals and wildlife. In the ecological toxicology field, biomarkers are generally defined as biochemical, physiological, or histological indicators of either exposure or effects of xenobiotic chemicals (Forbes et al., 2006; Adams and Rowland, 2003). As discussed in the Handbook of Toxicology: Indicators or biomarkers can be defined at any level of biological organization, including changes manifested as enzyme content or activity, DNA adducts, chromosomal aberrations, histopathological alterations, immune-system effects, reproductive effects, physiological effects, and fertility at the molecular and individual level, as well as size distributions, diversity in disease, and functional parameters at the population and ecosystem level.

Another definition of environmental biomarkers that has been used in ecotoxicology is that proposed by Shugart et al. (1992): “. . . a xenobiotically induced variation in cellular or biochemical components or processes, structures, or functions that is measurable in a biological system or sample.” Older biomarkers were focused on measures of whole animal physiology or biochemistry, whereas the newer biomarkers include measures of changes in gene expression. The utility of biomarkers in ecotoxicology is controversial, depending on the intended outcome of their use. In the ecological literature, ecological biomarkers of exposure used in environmental biomonitoring in the field are important for tracking substances and mixtures of concern (Hutchinson et al., 2005) and can provide a linkage between field and laboratory studies with a variety of organisms. Usage of these terms in ecological studies differs from the way these are used to report human public health, epidemiology, and biomonitoring trends. For additional information see also Kendall et al., 1998. Environmental biomarkers are expected to be mechanistically relevant, which may not necessarily be the basis for those used in public health, and reproducible (Hutchinson et al., 2005). The term biomarker has historically been used differently by field biologists and by laboratory scientists due to the desired outcome from their respective work. To accommodate both fields, many criteria have been developed to encompass all potential characteristics of a biomarker (for review see Forbes et al., 2006). The terms biomarker and biomonitoring differ

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within wildlife and ecology fields and are distinct from those used in human health disciplines so that no one definition is easily consistent across disciplines. This likely stems from a differential focus on individuals (humans) versus populations (wildlife). Furthermore, in the laboratory the term biomarker is rarely used, rather end point or clinical sign is more common. In addition, for many endocrine-relevant end points, the specific MOAs have not yet been identified, and hence these end points cannot be considered biomarkers of either exposure or effect. They are more general indicators possibly involving multiple MOAs in the endocrine system. Therefore, for the purpose of this chapter and the discussion on methods used to detect chemical exposures and effects, we have chosen to use end point as a general term to relay a measure of detection that is more quantitative (e.g., hormone concentration, fecundity). In this sense, an end point is any measurable parameter that is indicative of EAC exposure or effect. The term clinical sign is used to describe less quantitative clinical observations (e.g., head tilt).

18.3 END POINTS AND CLINICAL SIGNS ASSOCIATED WITH ENDOCRINE ACTIVITY End points available to assess EACs in the laboratory, field, and clinic vary across species and endocrine organ systems with respect to availability, accessibility, reliability, sensitivity, specificity, and general ease of use. Some such end points can be used to indicate changes in the endocrine system, whereas other end points are more general and not specific to the endocrine system but rather serve as indicators or clinical signs that can be linked to abnormalities of the endocrine system. Some general indicators in mammals of adverse effects from EACs include irregular cyclicity (in females), the formation of nonfunctional gametes (i.e., abnormal sperm), and infertility. Similar end points for endocrine activity in other taxa groups (i.e., fish, birds, amphibians, and invertebrates) include survival and the size of adults, fecundity (number of eggs spawned) or fertilization success (number of fertile eggs produced), hatchability or hatching success (number of fertile eggs that produce offspring), and offspring survivability (e.g., occurrence of malformations in hatched animals) (Hutchinson, 2004). For birds and fish, the sex ratio of offspring, the fertility of offspring at maturity, and the number of eggs produced by offspring can be indicative of altered reproductive viability of the offspring (Touart, 2004). Also, a number of end points (i.e., sex ratios, secondary sex characteristics, sexual behavior, and growth during early life stages) can be used across taxa (Hutchinson et al., 2005; Ankley and Johnson, 2004). Although some end points may differ between humans and other taxa, in this chapter, where possible, end points are compared across species to illustrate common markers of endocrine activity. Throughout this section, end points or studies that involve invertebrate species are described. The authors recognize that some biological systems in invertebrates are not conserved in vertebrates. However, many similarities have been established between invertebrates and vertebrates in the area of endocrinology such as developmental and neurohormonal effects and end points. Still, more work needs to be done on the extrapolation of endocrine effects from one species to another and certainly from invertebrate to vertebrate species, as well as the identification of critical differences between species that may make one species more sensitive to certain chemicals than others. At present, it is difficult to make definitive causal links between taxa. A chemical’s effects in one species are not always translatable to other species, and this is an area where more research and data collection across taxa would be helpful.

END POINTS AND CLINICAL SIGNS ASSOCIATED WITH ENDOCRINE ACTIVITY

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Some end points used to assess endocrine activity are general and can sometimes be caused by acting on a nonendocrine organ, leading to a downstream response by the endocrine system (i.e., an indirect effect). For this reason, general information is useful for population modeling; however, more diagnostic indicators of endocrine responses specific to MOAs can also accompany and hopefully be correlated with such general measures. Specific biomarkers of exposures to EACs include circulating levels of proteinaceous and steroid hormones, associated enzymes, and binding proteins. Also, the expression of secondary sex characteristics, histopathology of endocrine tissues, and vitellogenin (VTG) concentrations are considered to be specific end points that can be associated with exposure to EACs (Hutchinson et al., 2005). Some measurements for such end points can be accomplished with protein concentration measures and enzyme activity, while others are accomplished by measuring gene and protein expression. 18.3.1

Molecular, Cellular and Biochemical End Points

The following section describes examples of, but certainly not all, end points relevant to exposure to EACs in humans and other taxa groups. The end points discussed have been used to assess endocrine effects; however we emphasize that other end points of endocrine activity exist that are not listed here. A variety of molecular level assays exist to assess endocrine responses to chemical exposures (see Table 18.1). These assays often measure, for example, enzyme activities, mRNA synthesis, hormone concentrations, and changes in cellular structure. One enzyme class that provides a useful measure of function following chemical exposure is the hydroxysteroid dehydrogenases (HSDs), key enzymes in the steroidogenesis pathway that are involved in estradiol production. HSDs are good biomarkers because there is little natural variation in their activity levels (Kumar et al., 2000); however, different chemicals often result in a range of affinities for different HSDs and they may therefore be chemical specific (Rotchell and Ostrander, 2003). Another way to detect abnormal gene expression after exposure to chemicals is to measure mRNA production of key proteins. The mRNA synthesis of gonadotropins, which are derived from the pituitary gland and control gonad development and sex steroid synthesis, is disrupted by EACs, which in turn results in disrupted gonadal development (Harries et al., 2000). Like mRNA levels, changes in concentrations of key proteins within an organism can also be useful indicators of how a chemical affects homeostasis. Levels of neurohormones, such as vasotocin, gonadotropinreleasing hormone (GnRH), serotonin, corticotropin releasing factor (CRF), are relevant to effects following exposure to EACs. Some tests that are run routinely to assess thyroid activity in humans include tests for total and free thyroid hormones thyroxine (T4) and triiodothyronine (T3), and thyroid stimulating hormone (TSH) or thyrotropin; antithyroid peroxidase antibodies; and a marker for iodine status (e.g., urine iodine/creatinine). Altered steroid biosynthesis (e.g., aromatase) usually precedes histopathological changes and can therefore provide an early warning of endocrine effects (Rotchell and Ostrander, 2003; Guiguen et al., 1999). However, many controlled studies of exposure are needed to provide a range of compounds to build a profile of the catalytic properties for each biosynthesis enzyme for selected species of interest. Diagnostic tests that assess the hypothalamic–pituitary–gonadal axis in humans have varying degrees of sensitivity and specificity. Measurements of serum concentrations of follicle-stimulating hormone (FSH), luteinizing hormone (LH), free and total testosterone, and inhibin in men are used clinically, in part, to diagnose the basis of reduced spermatogenesis. In women, levels of sulfated metabolite of dehydroepiandrosterone (DHEA-S) are

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Major key hormones: progesterone, testosterone, estradiol Major key enzymes: HSDs, aromatise Thyroid hormone

C

C D

Major key hormones: T4 (thyroxine), T3 (triiodothyronine) in thyroid gland Major key enzyme: thyroid peroxidase Other hormone production measures: e.g., vitamin D in kidney Hormone Transport Serum/plasma hormone-binding proteins (e.g., alpha-fetoprotein; AFP and sex hormone binding globulin, SHBG) Thyroid hormone transport proteins: thyroxin-binding globulin (75%), transthyretin (15%), serum albumin Adrenal: corticosteroid-binding globulin (CBG), cortisol-binding protein (transcortin), serum albumin

Steroidogenesis

C

B

Neurohormones: vasotocin, GnRH, catecholamines, CRF, serotonin levels; ecdysteroid (molting hormone) receptor binding (invertebrates); ecdysone receptor (EcR) and juvenile growth hormone (invertebrates); Ala-Pro-Gly-Trp (ARGW) amide levels Pituitary Hormones: prolactin, LH, FSH, TSH, ACTH levels

Examples of Potential Endocrine End Points and Clinical Signs

Molecular and Cellular End Points

A, B

Letter Linked to Fig. 18.1

TABLE 18.1

(Fraser and Kodicek, 1970) (Davidson et al., 2006; OECD, 2006; Meulenberg, 2002)

(OECD, 2006; Bernal, 2005; Braverman et al., 2005; Eskiocak et al., 2005; Mukhi et al., 2005; Zaki et al., 2004; Gray et al., 2004; Laws et al., 2000; Stoker et al., 2000)

(Eskiocak et al., 2005; Zaki et al., 2004; Gray et al., 2004; Langer et al., 2003; Laws et al., 2000; Stoker et al., 2000) (Lavado et al., 2006a, 2006b; Ottinger et al., 2005; Gray et al., 2004; Rotchell and Ostrander, 2003; Odum and Ashby, 2002; McMaster et al., 2001; Kumar et al., 2000; Stoker et al., 2000; Gray, 1998; Soto et al., 1992)

(Ottinger et al., 2005; DeFur, 2004; Gagne and Blaise, 2003; Khan et al., 2001; Gao et al., 2000)

Reference(s)

669

Estrogen, progesterone, thyroid, androgens receptor binding Hormone Internalization and Action:

F

(Rotchell and Ostrander, 2003; Kester et al., 2002)

(Eidem et al., 2006; Ghekiere et al., 2006; Puinean and Rotchell, 2006; Nilsen et al., 2004; Tatarazako et al., 2004; Inui et al., 2003; Islinger et al., 2003; Rotchell and Ostrander, 2003; Hemmer et al., 2002; Oberdorster, 2001; Kanamori, 2000; Kumar et al., 2000; Panter et al., 1999; Celius and Walther, 1998; Arukwe et al., 1997; Flouriot et al., 1995)

(Gray et al., 2004; Vetillard et al., 2003; Bowman et al., 2002; Gray, 1998; Flouriot et al., 1995; Pakdel et al., 1991)

(Peeters et al., 2005; Fenske and Segner, 2004; Gray et al., 2004; Rotchell and Ostrander, 2003; Stoker et al., 2000)

The letters (A–H) in the table correspond to the letters (A–H) associated with the components of the endocrine system as illustrated in Fig. 18.1. The end points and clinical signs listed in the table are examples of end points that have been used in some endocrine assays; the list is not intended to be comprehensive or to include all end points of endocrine activity. This table combines human end points with those from other taxa groups.

G

Enzymes: deiodinases, aromatases, HSDs Hormone Recognition

F

VTG (vitellogenin, yolk protein) mRNA expression; VEP, ZR, and ZP, protein synthesis/detection; gcl and PCNA gene expression Metabolism/Metabolic Disruption: cytochrome P450s, glucoronidases, sulfotransferases, insulin, glucose, vitamin D metabolite (125s) levels

Hormone Conversion

E

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helpful, in addition to testosterone levels, for androgen testing that may be associated with infertility. Although commercially available hormone assay kits are relatively reliable, it is important to note the timing of the menstrual/estrous cycle when the test was done to aid in the interpretation of the results. The use of several valuable biomarkers has been demonstrated in taxa other than mammals. These include VTG, a major egg-yolk protein precursor produced by female oviparous vertebrates and some invertebrates, and vitelline envelope proteins (VEPs), glycoproteins that are components of the egg envelope that form the chorion, or the outer envelope surrounding the embryo, in the developing egg (Celius and Walther, 1998). VTG and VEPs are used to detect estrogenic and antiestrogenic exposures in fish, amphibians, birds and invertebrates (Hutchinson et al., 2005; Nilsen et al., 2004; Panter et al., 2002; Arukwe et al., 2000; Thorpe et al., 2000; Jobling et al., 1998; Matthiessen, 1998; Arukwe et al., 1997; Sumpter and Jobling, 1995). Also, the production of the chorion proteins zona radiata (ZR) and zona pellucida (ZP) is regulated by estradiol and can be more sensitive than VTG at low doses of estrogen exposure. (For a review of VTG, ZR, and aromatase in fish and invertebrates, see Rotchell and Ostrander, 2003.) Another protein that can be upregulated in response to EAC exposure is the estrogen receptor itself (Bowman et al., 2002; Pakdel et al., 1991). Commonly used end points for thyroid status are the thyroid hormones, T4 and T3, and the pituitary hormone, TSH (see Table 18.1). The majority of T4 is produced in the thyroid gland, but 80% of T3 is produced external to the thyroid through deionization of T4 (Sapin and Schlienger, 2003). In serum, the vast majority of T4 and T3 is protein bound (Sapin and Schlienger, 2003). Measurements for TSH receptor antibodies and for thyroid peroxidase and thyroglobulin antibodies are also available (Squire, 2006). In humans, individual variance in serum T4 and TSH is substantially narrower than the variance for the population (Andersen et al., 2002, 2003). Thyroid hormones are narrowly regulated around a “set point” that is predominantly controlled by genetics (Hansen et al., 2004). To illustrate, serum T4 and TSH are highly correlated between monozygotic twins (Hansen et al., 2004); in fact, to a significantly greater extent than between dizygotic twins. Population variability is a summation of individual variability. National estimates of the distribution of the biomarker end points of thyroid function have been published for several countries including the United States (Hollowell et al., 2002) and China (Teng et al., 2006); and Reference Intervals have been established for areas of Germany (Volzke et al., 2005) and Norway (Bjoro et al., 2000). National differences in distributions can reflect many factors including nutritional status (e. g., iodine intake) and potential exposure to environmental chemicals, as well as differences in genetic composition of populations. 18.3.2

Individual and Population-Level End Points and Clinical Signs

Individual and population-level end points and clinical signs used to assess endocrine activity range from general indicators (e.g., behavior and organ weights) to specific biomarker end points (e.g. secondary sex characteristics) (Table 18.2). Certain organs are individually examined and weighed following exposure to suspected EACs to identify abnormal increases or decreases in dry or wet weight as well as relative organ-to-body weight ratios (see Table 18.2). Increased weight of ventral prostate and seminal vesicles has been used to assess androgen exposure in rats (Gray et al., 2004). Many end points can elicit population-level effects as a result of exposure to EACs (see Fig. 18.2 as a case study of how end points associated with thyroid dysfunction can be

671

F

F F

F

F

F

F

F


TABLE 18.2

Histopathology: thyroid, brain, pituitary, adrenal glands, gross lesions, reproductive organs; and tissues include: testes, epididymes, seminal vesicles and coagulating glands, prostate, levator ani muscle plus the bulbocavernosus and glans penis, vagina, cervix, uterus, ovaries Mitotic index of epithelial lining of endometrium Target Organ Weight: organ-to-body weight ratio (i.e., GSI,gonadal somatic index; his, hepatic somatic index); thyroid weight/volume; organ-to-brain weight; absolute weight of organ(s) including: liver, kidneys, adrenal glands, pituitary, thyroid/parathyroid, ovaries, uterus, testes, seminal vesicle with coagulating glands, prostate, epididymis, levator ani and bulbocavernosus, Cowper’s gland, glans penis, brain, heart, spleen, and thymus Immunity: cytokine production, proliferation, T-cell function, induction of immunoglobulins, and autoantibody production by B1 cells Fertilization Success: fertility/infertility/# fertilized eggs/conception rates, embryo/larval/fetal/neonatal survival, # still and live births, time to brood release, hatchability (days to hatch), appearance of larvae, and premature labor Fecundity: spawning activity/frequency (group and pair breeding), number eggs produced Egg Integrity: shell thickness, # eggs cracked (birds) Growth: measured by body weight, hind limb length, body length, wing and bone length (birds), 28d growth LOEC (lowest observed effect concentration) (fish), strength Development: morphological development examples include: hypospadias, anogenital distance (AGD), undescended testes (cryptorchidism), urethral–vaginal distance, premature puberty; time to first foam, time to first egg, size cloacal gland/protuberance (birds), and molt frequency (invertebrates)

Examples of Potential Endocrine End Points and Clinical Signs

Whole Animal- and Population-Level End Points and Clinical Signs

(continued)

(Gray et al., 2004; Halldin et al., 2003; Laws et al., 2000; Gray, 1998)

(Kamata et al., 2006) (Mukhi et al., 2005; Gould et al., 1997)

(Thorpe et al., 2006; US EPA, 2002)

(Silva et al., 2005; Yurino et al., 2004; Nielsen and Hultman, 2002; Hultman et al., 1992) (Lawn et al., 2005; Ottinger et al., 2005; US EPA, 2002)

(Eskiocak et al., 2005; Tajtakova et al., 2005; Zaki et al., 2004; Gray et al., 2004; Langer et al., 2003; Rotchell and Ostrander, 2003; Laws et al., 2000; Stoker et al., 2000)

(Eskiocak et al., 2005; Zaki et al., 2004; Gray et al., 2004; Weber et al., 2002; Zillioux et al., 2001; Laws et al., 2000; Stoker et al., 2000)

Reference(s)

672

H

H

F. H

F, H

F, H

F


TABLE 18.2 Examples of Potential Endocrine End Points and Clinical Signs

Altered Reproductive Viability of Offspring: embryo/larval/juvenile viability, time to sexual maturation, fertility of offspring at maturity, number of eggs produced by offspring, sex ratio of offspring

Males: coloration/banding, nuptial tubercles, dorsal nape pad; preputial separation (PPS), nipple retention Females: size of ovipositor, urogenital papilla, vaginal opening, precocious breast development Male and female: plumage length, plumage dimorphism (birds); hirsutism (abnormal hair growth) (mammals) Sexual/Reproductive Behavior: territorial aggressiveness/defense, nesting (spawning), courtship, egg protection/care, nest attentiveness (birds), mounting, libido, erectile dysfunction (mammals) Formation of Functional Gametes: sperm number, motility, and morphology; gamete maturation (production, final oocytes maturation, sperm motility test) Sex Ratios: intersex, imposex, skewed sex ratios

Estrous/Menstral Cyclicity: irregular menses, amenorrhea, premature menopause Secondary Sex Characteristics:

(Continued)

(Maack and Segner, 2003; Orn et al., 2003; Rotchell and Ostrander, 2003; Seki et al., 2002; van Aerle et al., 2002; Oberdorster, 2001; Parks et al., 2001; Gronen et al., 1996, 1999) (Touart, 2004; Rotchell and Ostrander, 2003; Oberdorster, 2001)

(Nice, 2005; Horiguchi et al., 2002; Swan et al., 1997; Carlsen et al., 1992)

(Denslow and Larkin, 2006; Halldin et al., 2005; Ottinger et al., 2005; Oliva et al., 2002; Whelan et al., 1996)

(Gray et al., 2004; Rotchell and Ostrander, 2003; Quinn et al., 2002; Ankley et al., 2001; Harries et al., 2000; Laws et al., 2000; Stoker et al., 2000; Gray, 1998; Bortone and Davis, 1994; Bortone et al., 1989; Howell et al., 1980)

(McLachlan et al., 2006; Laws et al., 2000)

Reference(s)

END POINTS AND CLINICAL SIGNS ASSOCIATED WITH ENDOCRINE ACTIVITY

673

Tissue indicators Clinical signs or toxicological effects (histopathology)

Biochemical indicators

Molecular indicators

Polychlorinated biphenyls

None

Maternal T3, T4 with Infant TSH Anti-TPO and abnormal TSH (males only) Minor changes in tT4, fT4 and TSH

Thyroid gland volume Frequency of echogenicity

Inconsistent association between PCB exposure and thyroid hormone homeostasis

TH responsive gene expression of RC3/ Neurogranin TH responsive gene expression of myelin basic protein Serum tT4, fT4 and plasma tT4 Serum tT3, fT3 and plasma tT3

Colloid area

Hypothyroidism Neurological deficits Hearing loss

None

Plasma T4 (chicken embryos) Plasma T3 (mallards, kestrels) EROD (mallards) PROD (mallards) Hepatic type I monodeiodinases (chicken embryos) Thyroid gland weight (mallards)

Growth (femur length in chicken embryos)

FIGURE 18.2 Generalized effects of PCBs on thyroid end points.

indicative of various forms of clinical disease). Some such measures are visible only at the population level and not at the individual level (e.g., thyroid disease in human population as a result of iodine deficiency (WHO, 1997)) (see Table 18.2). These effects can bring serious long-term consequences to an entire population, although the individual health consequences are less well documented. Information on distribution of end points for the population remains an important contribution that generally has not been made. 18.3.3

Histology End Points

A variety of tissues are regularly examined using histology to detect potential pathological effects following exposure to EACs (see Table 18.2). The occurrence of testicular and ovarian tissue in the same gonad, apoptosis of gonad tissues, increased apoptosis of spermatocytes, Sertoli and Leydig cell hypertrophy or hyperplasia, stage or success of spermatogenesis/maturation of oocytes, status of seminiferous tubules, structural integrity and sperm production, and the number of uterine implantation sites/scars are all indications of altered reproductive tissues that can result from exposure to EACs. The thyroid is also often examined histologically in fish, birds, and laboratory rodents. Follicular cell hypertrophy or hyperplasia, characteristics of colloid, vascular supply, density/size/shape of thyroid follicles, and angiogenesis are all indicators of altered thyroid status. 18.3.4

End Points for HPT Effects

Size or volume of the thyroid gland and pathologies are often examined using ultrasound imaging (Wiesner et al., 2006; Braverman et al., 2005; Tajtakova et al., 2005; Langer et al.,

674


2003). Sonography is frequently used to determine thyroid volume in screening and is recommended (WHO, 1997) especially to assess iodine status of populations (Wiersinga et al., 2001; Brahmbhatt et al., 2000). Long-term iodine sufficiency influences thyroid volume creating difficulties in applying reference standards for screening with sonography developed on populations with marginal iodine sufficiency (WHO, 1997) to iodine-replete populations (Hess and Zimmermann, 2000; Foo et al., 1999; Xu et al., 1999). The same approaches for evaluating thyroid status are used in other species, especially dogs (Bromel et al., 2006; Reese et al., 2005; Peterson et al., 1997) and cats (Peterson, 2006; Daniel et al., 2002; Norsworthy et al., 2002; Mooney, 2001). Several examples are well described in literature documenting population-level effects of dietary components on endocrine status of domestic livestock and humans. The interrelationship between diet and contaminants is one that is often overlooked. Such examples may not readily come to the attention of persons accustomed to reading the environmental health literature. These are, nonetheless, important examples of population-level effects of a component of the environment (i.e., diet) on endocrine status. Although iodine deficiency is the most well-understood environmental cause of thyroid disease (Delange, 1994), worldwide additional food sources contain chemicals that produce clinically significant thyroid diseases including severe endemic goiter (Moreno-Reyes et al., 1993). Generally, the mechanisms of action appear to involve disruption of iodine utilization. Major examples of such antithyroid chemicals include hydrogen cyanide in cassava root (Teles, 2002), a major energy source for humans and domestic livestock that interferes with thyroid peroxidase activity; C-glycosylflavones from millet (Gaitan et al., 1989), a major energy source in semiarid tropical areas that inhibits thyroid peroxidases (Gaitan et al., 1989); and isoflavones from soy-containing foods and dietary supplements (Doerge and Chang, 2002; Doerge and Sheehan, 2002b). Although the mechanism of action of soy-containing foods in adversely affecting the thyroid is not fully understood, it may dispose individuals with iodine deficiency to hypothyroidism (Doerge and Sheehan, 2002b). Many additional plants contain chemicals having antithyroid activity (among others, see Chandra et al., 2004). Perchlorate, a chemical known to block the sodium iodide symporter (NIS) responsible for iodine uptake into the thyroid gland, has been found to be widespread in human milk that can serve as a biomarker for exposure to perchlorate (Kirk et al., 2005). Genetic defects in the iodination process of thyroid synthesis combined with environmental exposures and suboptimal iodine status have the potential to be clinically significant (Scinicariello et al., 2005). 18.3.5

End Points for HPG Effects

In species with prominent secondary sex characteristics (e.g., many fish species), skewed sex ratios can be easily and quickly identified at sexual maturation; however, identification of intersex gonads, such as the presence of testicular oocytes, requires histopathological analysis (Hutchinson et al., 2005; Ankley et al., 2003; Ankley et al., 2001; Harries et al., 2000). The early onset of puberty in humans can be considered a secondary sex characteristic end point of exposure. Observational studies in humans have identified end points as potential biomarkers of early developmental exposure and susceptibility to EACs (Ouyang et al., 2005; Colon et al., 2000). There is accumulating evidence that the early onset of puberty results in longer life exposure to estrogens and, thereby, may increase the risk of breast cancer (Lippman et al., 2001). Identifying such end points is critical in research with EACs in order to provide linkage to population-level relevant effects. Longitudinal studies

ENVIRONMENTAL CHEMICALS AND END POINTS: CASE EXAMPLES

675

from early childhood to adulthood are required to identify clear biomarker end points of effect regarding the risk associated with developmental estrogen exposure, breast cancer in women, and, perhaps, other developmental malformations and cancers in both women and men. 18.3.6

End Point and Clinical Signs Summary

As discussed above, several end points and clinical signs exist, ranging from behavior and molecular indices in individuals to population-level markers that can indicate effects from exposure to EACs. Some of the end points and clinical signs discussed above and mentioned in Tables 18.1 and 18.2 are of high ecological importance (i.e., fecundity, fertility, altered sexual behavior, secondary sex characteristics, and sex ratio). However, many of these end points and signs can be affected by changes in the environment (e.g., temperature, food availability, age, and reproductive status), including, but not limited to, exposure to contaminants, and hence do not offer much information on the potential MOA of the chemicals of interest (Denslow and Larkin, 2006). If these end points are utilized in concert with mechanistically relevant assays, they can elucidate more information about MOA or pathway of such effects, while still making the link to the population level. Molecular and cellular in vitro assays often offer more sensitive end points and may, therefore, be able to provide early warning of adverse effects in individuals and at much lower chemical concentrations than in vivo approaches (Denslow and Larkin, 2006). The usefulness of molecular markers greatly enhences if they can be used to make the link to adverse effects on the individual and/or population. Recently, efforts have been made using “systems toxicology” (Waters and Fostel, 2004) to describe changes in gene expression to toxicants from a whole animal-level approach. Also, “phenotypic anchoring” (Tennant, 2002) stresses the correlation of changes in gene expression with traditional end points that lead to adverse effects. These approaches are notable, but still more research is needed to create linkages between molecular approaches and adverse individual and population effects. Clinical signs that are currently related to endocrine effects in humans as a result of exposure to potential EACs are useful, but continued research is needed to develop more sensitive end points that can serve as relevant biomarkers to link the exposure of potential EACs to endocrine disruption. In this regard, value and insight have been and will continue to be gained in studying various taxa in the laboratory and ecoenvironment. The hope is that the development of better technological tools will result in more specific end points that can be used as biomarkers to help in diagnosis and treatment and, ultimately, in the prevention of overexposure to potential EACs that may result in adverse effects on human and animal individuals and populations.

18.4 ENVIRONMENTAL CHEMICALS AND END POINTS: CASE EXAMPLES Case 1. Thyroid Hormone and Three Environmental Chemicals with Different Mechanisms of Action The endocrine actions of thyroid hormone can be disrupted at multiple levels such as synthesis, iodination, release of stored hormone, transport by plasma proteins, peripheral activation, target organ metabolism (e.g., brain, liver), deiodination, and excretion of

676


breakdown products. Recently, the role of the deiodinases, in local control of thyroid hormone at the tissue level, and implications for altering whole animal physiological status have been more thoroughly described (reviewed by Bianco and Kim, 2006). Environmental chemicals can affect multiple and different steps in this process. A description of three chemicals that affect thyroid hormone synthesis, metabolism, and excretion is provided in Fig. 18.2a–c. The chemicals, lithium, perchlorate, and polychlorinated biphenyls (PCBs), chosen have distinctly divergent MOAs on thyroid hormone as well as different environmental characteristics. Lithium. Lithium, the so-called ‘hard’ ion that binds to oxygen-donors better than do the hard metal ions, which are utilized biologically (Na, K, Mg, and Ca) is in the ‘a’ class of hard metals (Frau´sto da Silva and Williams, 1993). Lithium is perhaps best known for its pharmaceutical use, especially to treat psychiatric conditions. Data from the National Health and Nutrition Examination Survey (1999 through 2001) have provided population estimates for the United States that indicate pharmaceuticals containing lithium were taken by approximately 0.1% of the population, representing around 220 000 people in the United States (Aoki et al., 2006). Lithium is found at low concentrations in the major rivers in the United States (Kszos Lakowicz and Stewart, 2003). Although best known for pharmacological properties, lithium has a number of industrial uses (see Moore, 1995 for summary) including as a component of batteries; research is going on to produce nanoparticles for improved lithium ion battery cathodes (Liu et al., 2006). It is also used as a catalyst in chemical reactors and as a cell additive in aluminum production. Developmental and reproductive toxicity of lithium to mammalian species is well established. Moore and the Institute for Evaluating Health Risks Expert Scientific Committee (IEHRSC) assessed the reproductive and developmental toxicology literature for lithium utilizing an expert evaluative process concluding that the data were sufficient to indicate that lithium at therapeutic doses can cause developmental toxicity in humans and that lithium causes developmental toxicity in both rats and mice based on data from studies using both prenatally and postnatally dosed animals (based on studies that included the period of lactation to weaning and beyond) (Moore, 1995). The teratogenic effects of lithium on nonmammalian species are so pronounced that lithium salts are used as a laboratory “standard” to create teratogenicity in frog species. As evaluated by Moore and the IEHRSC, some of the reproductive studies in rodents observed changes with lithium dosing near clinical therapeutic concentrations, although these studies were not used for quantitative risk assessment. This work did not include avian, amphibian, or fish species and noted that no data specifically on mammalian male reproductive toxicity in the literature were found (Moore, 1995). In a more recent study in nonrodent species (Banerji et al., 2001), specifically birds (i.e., finches), lithium treatment decreased testicular weight and seminiferous tubule diameter and induced severe degenerative changes in germ cells at serum lithium concentrations in the therapeutic range of plasma lithium for humans (0.5 to 1.5 mEq/L or 3.5 mg/L to 10.5 mg/L). Although comparison of responses to lithium across species, as illustrated in Fig. 18.2a, indicates several similarities with respect to thyroid, the reproductive tracts of male birds appear to be especially sensitive to the effects of lithium exposure. Clearly, the sensitivity of different tissue types to the effects of lithium differs across species. Whether concentrations of lithium producing teratogenic or reproductive effects are obtained in nature is a wider topic than can be addressed in this brief review. Background concentrations of lithium in soil average 30 mg/kg to which lithium from


677

sources such as volcanic activity, fossil fuels, and disposal of lithium-based batteries are added (Rydh and Svard, 2003). The effects identified in the nonmammalian species are severe. Exposures that produce subtle, adverse effects appear to be unknown. Moore and the IEHRSC recommended that studies be undertaken on developmental toxicology that included long-term assessment of kidney, heart, thyroid, and central nervous system (CNS) structure and function (Moore, 1995). Data are available to observe some of the effects of lithium on thyroid in multiple species, including nonrodent species. Lithium also produces a variety of adverse effects on the that are summarized in Fig. 18.2a for humans (Bocchetta and Loviselli, 2006; Lazarus, 1998; Loviselli et al., 1997; Bocchetta et al., 1996; Mizukami et al., 1995; Perrild et al., 1990; Bagchi et al., 1978; Rosser, 1976); rats (Allagui et al., 2006; Frankenfeld et al., 2002; Chanoine et al., 1993; Chatterjee et al., 1990; Etling et al., 1987; Bagchi et al., 1978; Berens et al., 1970); and birds (Downie et al., 1977). Perchlorates. Perchlorates are widely distributed environmental contaminants, not bioaccumulative, and are also used as drugs. Perchlorates form naturally in the atmosphere and are present in precipitation (Dasgupta et al., 2005), concentrate in particular regions (Dasgupta et al., 2005), and can be greatly increased through use of various products including explosives, pyrotechnics, and solid rocket propellant (Miediratta et al., 1996). Blount and colleagues reported that among women who had been participants during the years 2001 and 2002 in the National Health and Nutrition Examination Survey in the United States whose urinary iodine was less than 100 mg/L (but not those whose urinary iodine was more than 100 mg/L), increased perchlorate was associated with lower production of T4 and an increase in TSH to stimulate additional T4 production (Blount et al., 2006). This is the first time that effects of environmental levels of perchlorate have been associated with thyroid hormone decrements in a general population study at ambient contamination levels. The general trends observed in the perchlorate literature for humans, rodents, amphibians, and birds are represented in Fig. 18.2b. These studies demonstrate generalized trends across taxonomic groups for the thyroid hormones and thyroid gland effects. In human females and all rodents, serum T4 declines and TSH increases following exposure to perchlorate, while in birds T4 and T3 were shown to decline and in frogs only T3 was demonstrated to decline. Similarly, thyroid gland weight or volume increased in human, rodent and bird examples. Many of the effects listed are taxa specific and represent a limited number of studies. Therefore, it is important to note that the effects outlined in Fig. 18.2 demonstrate generalized trends seen in the literature. These figures are intended to demonstrate the importance of observing effects across species, demonstrating the value of a weight-of-evidence approach for use in risk assessment. The studies utilized in Fig. 18.2 represent general trends observed and do not demonstrate the effects of each study in the literature due to differences in experimental design such as exposure times, exposure duration, chemical administration, and sex and age of the animal studied. Studies summarized in Fig. 18.2b include those for humans (Blount et al., 2006; Kirk et al., 2005; Greer et al., 2002), rodents (OECD, 2006), birds (McNabb et al., 2004a, 2004b), and frogs (Zhang et al., 2006; Tietge et al., 2005; Fort et al., 2000). Polychlorinated Biphenyls. The PCBs make up a mixture of different biphenyl compounds that can be chlorinated in different ways to include 209 separate chemical structures or congeners. The PCBs were designed for use in transformers, as lubricants in electrical equipment, as plasticizers, in surface coatings and inks, in carbonless duplicating paper, and as an extender in pesticide mixtures (http://www.epa.gov/ttn/atw/hlthef/polychlo.html).

678


The PCBs are highly lipid soluble, bioaccumulative, and persistent and continue to exist in the environment decades after their use was discontinued in the 1970s. Commercial production of the PCBs was banned in 1979 due to their ability to accumulate in the environment (Ross, 2004; Carpenter, 1998). These congeners have been detected in many environmental media including air, water, aquatic and marine sediments, fish, wildlife, and human adipose tissue, serum, and milk (Safe, 1993). It is thought that exposures were highest in the late 1970s and early 1980s, declining after manufacturing ceased and regulatory controls were enacted (Ross, 2004). As a chemical class, the PCBs have multiple toxic effects and are well studied as antithyroidal agents (Carpenter, 1998). Various PCB congeners are known to interact directly with the thyroid gland and alter the ability of T4 to bind to the serum binding protein, transthyretin (TTR), and they have also been shown to alter the expression of liver glucuronidase, and thus the normal metabolism of the thyroid hormones (OECD, 2006). Exposure of rats to PCBs can produce thyroid changes. The National Toxicology Program bioassay of PCB 153 across a wide range of doses showed mild follicular cell hypertrophy at the highest dose. PCB 153 caused nonneoplastic lesions on the thyroid gland in female rats (NTP, 2006). Similar results are reported for PCB 126, and data support a MOA for PCB 126 involving induction of hepatic uridine diphosphate glucuronyl transferases activity (Fisher et al., 2006). In other nonrodent species, increasing PCB concentrations in blubber of harbor seals is inversely associated with serum total T4, free T4, total T3, and free T3, as well as seal thyroid hormone receptor (TR) activity (Tabuchi et al., 2006). PCBs are known to affect brain development through their effects on thyroid (reviewed by Koibuchi and Iwasaki, 2006; Roegge and Schantz, 2006; Zoeller et al., 2000), although the mechanism is not well characterized. In humans, PCBs reduce circulating T4 concentrations that could selectively damage the cerebellum during in utero development (potential mechanisms reviewed by Roegge and Schantz, 2006). Exposure to PCBs has been associated with lower levels of total T4 and free T4 in women (Persky et al., 2001) and a significant negative association between T3, T4, and TSH with PCBs has been reported for adult men (Turyk et al., 2006). These associations are, however, not consistent across populations (Hagmar et al., 2001a; Hagmar et al., 2001b; Osius et al., 1999). The generalized effects of PCBs on the thyroid systems of humans, rodents, and birds are summarized in Table 18.2c. Figure 18.2c shows a general trend across species of a decline in thyroid hormone levels following exposure to PCBs. Specific effects are noted for each taxonomic group based on literature reviews for humans, rodents, and birds. The effects represented in Fig. 18.2c show general trends summarized in the literature, but by no means represent either every study and exposure scenario or every congener: humans (Langer et al., 1998, 2005; Hagmar, 2003; Longnecker et al., 2000; Koopman-Esseboom et al., 1994); rodents (OECD, 2006; Zoeller et al., 2000; Brouwer et al., 1998; Goldey et al., 1995); and birds (Smits et al., 2002; Gould et al., 1999; Fowles et al., 1997). Observing the three examples presented in Fig. 18.2a–c, it is clear that these three, highly evaluated chemicals can disrupt thyroid hormone function at multiple and diverse steps in the synthesis and metabolic processes. Likewise, information obtained by using different taxa is complimentary, providing useful information in that responses not identified in one species may be apparent in another. Relying on in vitro assays or in vivo assays, in isolation, using single species, can produce only a partial picture of the endocrine activity of a chemical. Likewise, it is essential to note species differences (e.g., see Fig. 18.2a–c), so that these differences can be interpreted to most accurately assess the effects of a chemical on the environment.


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Case 2. Lead and the Vitamin D/Endocrine System Much interest, research, and programmatic activity have focused on EACs that disturb the estrogenic, androgenic, and thyroid systems. This reflects the origins of the work in reproductive toxicology and observations among wildlife species (among others see EPA, 1998; Reiter et al., 1998; Cooper and Kavlock, 1997; Colborn et al., 1993). Major, critically important areas of endocrine function have largely been omitted from consideration, including, for example, the adrenals (Harvey and Everett, 2003), the pancreas, and other systems such as the vitamin D/endocrine system. Adverse health outcomes following exposures to inorganic lead have been very well documented as an environmental health problem over the past 40 years. Risk assessments for inorganic lead have addressed the following organ systems: hematopoiesis, the kidney, and the nervous system (see chapter by Mahaffey et al., 2000). Endocrine effects of inorganic lead exposures have never been routinely explored, so have never been a central part of risk assessments. The example that follows describes how inorganic lead affects the vitamin D/ endocrine system and refers to the steps described in Fig. 18.1. Vitamin D is a prohormone produced in the skin through ultraviolet irradiation of 7-dihydrocholesterol (DeLuca, 2004), as well as being a provisionally required nutrient. More than 90% of the vitamin D requirement for most people is met by exposure to sunlight rather than diet (Holick, 2004). The physiological role of vitamin D in regulating calcium and phosphorus metabolism received the earliest attention following observations that rickets was associated with lack of exposure to sunlight (reviewed by Holick, 2004). With advancing experimental methods, elucidation of the cellular biology of the vitamin D receptor and genetic control of the vitamin D–endocrine pathways has advanced and been the topic of numerous reviews (among others, see Kiraly et al., 2006; Dusso et al., 2005; DeLuca, 2004; Holick, 2004; Prosser and Jones, 2004; Fleet, 2004). The biochemical and physiological changes caused by the vitamin D/endocrine system have been reviewed for multiple organ systems asuch as the central nervous system (Kiraly et al., 2006), the kidney (Dusso et al., 2005), and the immune system (DeLuca, 2004; Holick, 2004), as well as the pathogenesis of diabetes (Reis et al., 2005; Mathieu and Badenhoop, 2005; Ogunkolade et al., 2002). Hormone Production. One of the toxic effects of inorganic lead is interference with the production of 1, 25-dihydroxyvitamin D, which is, in turn, known to affect the storage and release of lead (Onalaja and Claudio, 2000)(see Fig. 18.1, Box C). Hormonal vitamin D (i.e., 1,25-dihydroxyvitamin D) interacts with nuclear vitamin D receptor (VDR), an ancient member of nuclear receptors for steroid hormones (Dusso et al., 2005), regulating the production of calcium-binding proteins (Onalaja and Claudio, 2000). Decreased production of the metabolically active form (1,25-dihydroxy-vitamin D) due to lead exposure affects the amount of hormonal vitamin D available to bind and activate the VDR, and thus ultimately affecting production of calcium-binding proteins. Hormone Conversion. Preformed vitamin D2 and vitamin D3 are obtained directly from the diet. 7-dehydrocholesterol (present in skin) can be converted into vitamin D3 by UV sunlight (Holick et al., 1977). These preformed chemicals are metabolically converted into hormonally active chemicals in the skin, known as prohormones (see Fig. 18.1). These prohormones are then metabolized in the liver into the steroid hormone, 25-hydroxyvitamin D3. The hepatic cytochrome P-450 CYP2R1 appears to be the critical 25-hydroxylase involved in vitamin D metabolism (reviewed by Dusso et al., 2005). The second step of vitamin D activation is conversion of 25-hydroxyvitamin D to 1,25-dihydroxyvitamin D,

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occurring under physiological conditions primarily in the kidney (Fraser and Kodicek, 1970). Fig. 18.1 describes these steps as hormone conversion (Fig. 18.1, Box E). Inorganic lead interferes with the production of the metabolically active form of vitamin D. Reduced plasma 1,25-dihydroxyvitamin D occurred in lead-fed rats (Smith et al., 1981). Lead-fed chicks had reduced conversion of 25-hydroxyvitamin D to 1,25-dihydroxyvitamin D (Edelstein et al., 1984). In epidemiological studies children with blood lead levels varying from 12 to 120 mg Pb/dL, serum concentrations of the vitamin D hormone, 1,25-dihydroxyvitamin D, were strongly correlated (r ¼ 0.88) and reduced to levels found in patients with metabolic bone disease (Mahaffey et al., 1982). In lead-poisoned children, 1,25dihydroxyvitamin D concentrations returned to normal shortly after chelation therapy, whereas, serum 25-hydroxy vitamin D levels did not change with treatment (Rosen et al., 1980). Hormone Recognition, Internalization, and Action. All biological responses of vitamin D arise following its metabolism into secosteroids with two principal metabolically active forms: 1, 25-a-dihydroxyvitamin D3 and 24R, 25-dihydroxyvitamin D3. With regard to Fig. 18.1, these processes would be included as part of metabolism and elimination (Fig. 18.1, Box G). 1,25 –a-dihydroxyvitamin D3 is the dominant metabolite producing a wide array of biological responses by interacting with the vitamin D nuclear receptor that regulates gene transcription in over 30 target organs with a cell membrane receptor that mediates rapid biological responses (Norman et al., 2002). Vitamin D metabolites are best known as regulators of ionized calcium homeostasis through actions on the intestines, the kidneys and bone, collectively called the vitamin D endocrine system (MacDonald et al., 1994). In the schematic shown in Fig. 18.1, these steps would be in the area of hormone recognition, internalisation, and action (Fig. 18.1, Boxes F and G). In addition to affecting calcium metabolism, vitamin D exerts a number of nonclassical effects (as reviewed by Dusso et al., 2005): suppression of cell growth, regulation of apoptosis, modulation of immune response, control of keratinocyte differentiation and function in the skin, and control of other organ systems including the renin– angiotensin system, insulin secretion, muscle function, and neurological function. Whether or not there are effects of lead on metabolism and elimination of the metabolically active form of vitamin D, as shown in Fig. 18.1 (Box G), has not been entirely clarified; however, because exposure to inorganic lead can interfere with conversion of vitamin D to its active metabolite, it can be surmised that this effect could ultimately impact the function of the organ systems reliant upon active form of vitamin D. Investigations into this form of endocrine disruption are warranted and should potentially be considered for risk assessment. Latent and Population-Level Effects. The last step identified in Fig. 18.1 addresses latent and population-level effects of the interaction of the environmental chemical and the endocrine system (Fig. 18.1, Box H). Broadly speaking interactions between environmental agents, diet, genes, and disease are areas of great interest in advancing disease prevention (Willett, 2002). Combinations of such factors are known to increase risk of disease. Interactions between lead and vitamin D, as previously described, have been investigated experimentally over several decades. In recent years, it has become clear that there are genetic differences in the population distribution of various forms of the receptors and in the biology of the various vitamin D receptor polymorphisms (Fleet, 2004; Uitterlinden et al., 2004). These, as well as the nature of human microsomal vitamin D3 -25-hydroxylase remain areas of active investigation. Vitamin D receptors are anticipated to affect lead storage and/or release (Onalaja and Claudio, 2000). The influence of VDR gene polymorphisms on VDR protein function and

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signaling is largely unknown (Uitterlinden et al., 2004). These differences could help assess the population-level impact (see Fig. 18.1, Box H) of interactions between exposure to inorganic lead and vitamin D effects and emphasize that there may be sensitive populations. To illustrate population-level effects, the role of genetic differences has been studied. Three vitamin D receptor gene polymorphisms (BB, tt, and AA genotypes) were associated with low mineral density (Fountas et al., 1999). One of the best-characterized restriction fragment length polymorphisms, Bsml (Uitterlinden et al., 2004) has also been studied with lead (Onalaja and Claudio, 2000). Schwartz and colleagues found that among former organolead workers, bone lead was higher in workers with the BB genotype, intermediate in the heterozygous, and lower in the bb homozygous workers (Schwartz et al., 2000). However, results across studies have been variable. Recent investigations of modifications by polymorphisms in genes encoding the VDR and the association between lead exposure and biomarkers of lead exposure and effect have either found no association (Weaver et al., 2003) or found that a particular polymorphism of the VDR gene (VDR B allele rather than the VDR bb genotype) was associated with high lead burden (Theppeang et al., 2004) and that renal outcomes were more adverse in younger lead workers with the variant VDR B allele (Weaver et al., 2006). Among children, the VDR gene polymorphism VDR-Fok1 has been found to modify the association between exposure to inorganic lead and blood lead concentrations based on observations in young children 6–24 months of age (Haynes et al., 2003). Children with the FF genotype had a greater increase in blood lead concentration than did children with the Ff genotype. Based on a different VDR marker (VDRBB and VDRBb), Shi and colleagues reported that 5–6-year-old children in a Chinese kindergarten who had the VDR B allele reported significantly higher blood lead levels than those with the VDR bb genotype (Shi et al., 2003). Overall, at the latent or population level (see Fig. 18.1, Box H) the associations between vitamin D receptor gene polymorphisms and bone disease are complex. A recent metaanalysis of data from more than 26,000 participants involving 9 European research teams concluded that the FokI, BsmI, ApaI, and TapI vitamin D receptor polymorphisms are not associated with metabolic bone disease or fractures, but the Cdx2 polymorphism may be associated with the risk of vertebral fracture (Uitterlinden et al., 2006). Similarly, MacDonald and colleagues found no significant association between common polymorphisms for the VDR and bone mass, bone loss, or fracture (Macdonald et al., 2006). In view of the limited impact of lead on bone mineral through vitamin D receptor, identifying an effect of lead on bone density or bone mass would not be expected to raise substantial questions about the population-level impact of this association.

18.5 DEVELOPMENTAL ORIGINS OF HEALTH AND DISEASE After describing some of the end points that are indicative of endocrine activity and providing an example of how a widely distributed environmental contaminant, inorganic lead, affects a relatively less well-known endocrine system, we now turn to two important areas that are of comparatively recent interest. A research area that has become a major focus is the study of how EACs may act through the concept of Developmental Origins of Health and Disease (DOHaD). This is the most recent terminology used to describe how the onset of adult diseases may be induced in utero/in ovo during embryo/fetal development, remain latent during neonatal and pubertal development, and emerge as an acute or chronic disease

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in adulthood. In the late 1980s, the “Barker hypothesis” was introduced, which was primarily based on human epidemiological data relating low birth size and poor maternal nutrition during fetal development to an increased risk of noncommunicable diseases (e.g., coronary heart disease, type-2 diabetes, osteoporosis, and metabolic and endocrine dysfunction) in the adult offspring (Barker, 2003). This led to the development of the paradigm, “Fetal Origins of Adult Disease” or “Fetal Basis of Adult Disease” (FeBAD) (Lau and Rogers, 2004). The paradigm was initially criticized, but has since gained support. Now FeBAD no longer remains restricted to just maternal nutritional effects on the fetus but also applies to chemical exposure and to other life stages, pre- and postnatally. For example, in animal studies, the effects of maternal exposure to environmental toxicants during pregnancy have led to permanent morphological and physiological changes in the developing embryo and neonate that apparently predispose the F1 offspring to diseases later in adult life (Lau and Rogers, 2004), some of which are similar to those observed in adult humans as discussed in the next section on transgenerational effects, which do not necessarily involve in utero exposure. The role of EACs in DOHaD is difficult to decipher in humans due, in part, not only to moral and ethical constraints but also to the latency of effects and complexity and scope that the endocrine system has alone and in concert with other major physiological systems (e.g., neurological and immunological). Nonetheless, a classic example of environmental endocrine exposure in utero and the dire consequences in the F1 offspring thereafter is the effect of diethylstilbestrol (DES). A historical account of the development and use of DES as an alternative therapeutic source of estrogen for women during pregnancy has been reviewed (Newbold and Jefferson, 2006). Briefly, a clinical report in 1971 indicated that a rare form of reproductive tract cancer, vaginal adenocarcinoma, was detected in adolescent daughters of women who had taken the drug while pregnant. Although the US Food and Drug Administration immediately restricted the use of DES during pregnancy, it has since been observed that daughters of mothers who had taken the drug while pregnant have had various other abnormalities, including reproductive organ dysfunction, abnormal pregnancies, decreased fertility, immune system disorders, and behavioral problems. In addition, DES sons were found to have structural and functional abnormalities, including hypospadias, microphallus, retained testes, inflammation, and decreased fertility. Notably, the deleterious effects of DES have been related to the time of exposure in utero (i.e., more adverse outcomes are associated with exposure earlier during gestation than later). Many of these developmental abnormalities observed in humans have been observed in mice following prenatal exposure to experimental doses of DES (Newbold et al., 2006). Rodent studies clearly suggest that DES exposure to pregnant dams increases susceptibility of the F1 offspring to malignant tumors of the reproductive tract (for review, see Newbold and Jefferson, 2006). In addition, numerous studies in pregnant rodents have indicated an increased incidence of mammary tumors (Fenton, 2006; Walker, 1992; Rothschild et al., 1987) and increased prostate-specific androgen receptor binding and organ weight when respective F1 offspring of DES-exposed dams reached adulthood (Gupta, 2000; vom Saal et al., 1997). It is equivocal, however, whether breast tumor development in daughters and prostate tumor development in sons of DES-exposed mothers are comparable to that observed in laboratory animals. Apart from DES, no other estrogenic chemicals have been documented to be directly linked to adult diseases in humans. However, there are numerous studies with pregnant animals that have shown in utero exposure of the embryo/fetus to synthetic and natural environmental compounds with estrogen-like activity (e.g., bisphenol A, atrazine, dioxin, and genistein) accelerates mammary gland development (Birnbaum and Fenton, 2003) and

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prostate size in the F1 offspring. In this regard, accentuated growth may lead to an increased susceptibility of the adult animal to chemicals with estrogenic activity and, perhaps, a predisposition to breast and prostate cancers. For example, in female rats, prenatal or perinatal exposure through mothers (or externally during the perinatal period) to bisphenol A, a chemical compound used in epoxy resins and polycarbonate plastics, stimulated early mammary gland development in the F1 offspring that was associated with an increased risk of mammary gland tumor development as the animals reached adulthood (Munoz-de-Toro et al., 2005; Markey et al., 2001; Colerangle and Roy, 1997). In male rats, exposure of the pregnant mother to bisphenol Awas associated with an increased sensitivity to estrogen and increased prostate weight in the male F1 offspring with age. Dioxin, an industrial waste byproduct, and genistein, a phytoestrogen found in soy-based products, increased the susceptibility of exposed rodent F1 offspring to carcinogens in adulthood (Birnbaum and Fenton, 2003). Perinatal exposure to atrazine, a herbicide, has been associated with increased prostate gland inflammation in adult rats (Stoker et al., 1999). Atrazine and dioxin exposure in utero delayed mammary gland development in early postnatal life, as early as postnatal day (PND) 4 in rat F1 offspring. Both compounds alter development of the neonatal mammary fat pad. Thus, prolonging maturation of the mammary gland during adolescence lengthens the period in which mammary tissue is vulnerable to potential carcinogens that may result in mammary gland tumors in adults (Birnbaum and Fenton, 2003). More recently, exposure of pregnant rats to experimental doses of vinclozolin or methoxychlor during the period of embryo gonadal differentiation in utero resulted in decreased spermatogenic capacity and increased male infertility in F1 and subsequent generations of adults (Anway et al., 2006). The concept of DOHaD has also been documented in nonmammalian species. For example, latent effects of EACs have been repeatedly demonstrated in Japanese quail (Coturnix japonica). A single in ovo injection of a relatively low dose of DES before sexual differentiation altered reproductive behavior of F1 adult males (Vigiletti-Panzica et al., 2005; Halldin et al., 2005). Elements of copulatory behavior necessary for successful mating that include neck grab, mount attempts, successful mounts, and cloacal contact movements were significantly altered; however, gonadosomatic index, testes weight asymmetry, serum testosterone concentration, and cloacal gland area were not significantly affected (Halldin et al., 2005). In a comparable study with a higher dose of DES administered in ovo prior to sexual differentiation (Vigiletti-Panzica et al., 2005), male copulatory behavior of the F1 adult was again altered and was also associated with a significant decrease in density of vasotocin immunoreactivity in the medial preoptic nucleus, bed nucleus of the stria terminalis and lateral septum of the central nervous system (CNS). The results of these studies demonstrate that direct in ovo exposure to an estrogenic compound can exert permanent effects on sexually dimorphic regions of the CNS that are not expressed until the individual reaches adulthood. Despite a wealth of observational or epidemiological data on humans and empirical information on various mammalian and nonmammalian animal models, a direct link between in utero/ovo exposure to EACs and potential adverse effects in the F1 generation of adults has not been established, apart from the effects of DES. More research is necessary to understand the mechanisms (e.g., genomic and epigenomic) and effects of exposure at various developmental life stages (e.g., embryo, fetal, and neonatal) as well as the influence of various lifestyles (e.g., nutrition, physical activity, occupation, and addictions) between the time of exposure during pregnancy and the time at which disease, initially induced in utero/ovo, is expressed in the F1 generation of adults.

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18.6 TRANSGENERATIONAL EFFECTS To bring this chapter to a close, we now turn to another research area of substantial interest in understanding the mechanisms of action of EACs, specifically transgenerational effects. These effects are defined differently within various groups among the scientific community. Some scientists define a transgenerational effect as one that is simply transferred from the parent generation to the next or first (F1) generation of offspring (Campbell and Perkins, 1988), whereas others define transgenerational effects as having been elicited minimally through the third (F3) generation (Anway and Skinner, 2006), which would likely be due to a genetic (DNA) or an epigenetic (e.g., DNA methylation) alteration. According to the first definition, trangenerational effects may be applicable to the concept of DOHaD via maternal transfer. For the purposes of this review, we do not differentiate between these two definitions. As was described in the previous section, exposure of pregnant mothers to DES is capable of causing latent developmental effects on the F1 generation following in utero or in ovo exposure. These latent effects of DES, however, may be distinct from the concept of DOHaD since F2 and, perhaps, F3 generations of sons from grandmothers initially exposed to DES have been observed with an increased risk of hypospadias (Klip et al., 2002). This study should be viewed with caution, however, as only a small number of affected grandsons were observed (Hernandez-Diaz, 2002). Nonetheless, researchers continue to study subsequent generations of sons and daughters whose grandparents were initially exposed to DES to determine if a transgenerational MOA of DES occurs in humans. Rodent studies have also reported DES alterations to germ cells that were then passed on to later generations of offspring who were never directly exposed to DES (for review see Newbold and Jefferson, 2006). Distinct from the concept of DOHaD, trangenerational effects of EACs have also been observed in nonmammalian species, specifically in fish exposed to DES prior to spawning. In these studies, the F1 generation was not exposed directly to DES, but they experienced effects due to their parent’s exposure before spawning. In one of these studies, a fish lifecycle test conducted on the Chinese rare minnow (Gobiocypris rarus) observed transgenerational effects of DES on the F1 generation, including decline in the survival of F1 fry, decrease in testosterone in male progeny, and a slight elevation of estrogen levels in female progeny (Zhong et al., 2005). Other estrogenic chemicals have been demonstrated to have similar transgenerational effects on fish. In another study, Schwaiger and colleagues showed that the surfactant nonylphenol, an alkylphenol with weak estrogenic activity, leads to hormonal imbalances in the offspring of adult rainbow trout (Oncorhynchus mykiss) exposed to DES before spawning (Schwaiger et al., 2002). F1 generation offspring in this study had increased plasma levels of vitellogenin, estradiol, and testosterone in females and increased plasma estradiol in males. The fish data demonstrate that transgenerational effects occur in multiple taxa, and this phenomenon is thus not exclusive to any one species. The mechanisms required to induce transgenerational effects have not been thoroughly elucidated, and this is currently an exciting area of research. One mechanism described that could lead to transferable effects across multiple generations is epigenetic inheritance, also referred to as imprinting. Although epigenetic inheritance has not been investigated in most of the transgenerational studies in literature, several groups have begun to demonstrate epigenetic changes, specifically in methylation patterns in germ cell lines, which could lead to changes in exposed individuals that would be passed to their offspring (Anway et al., 2005, 2006; Anway and Skinner, 2006). Some of the chemicals studied

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include DES, vinclozolin, methoxychlor, and Cr (III) chloride. These studies are described in more detail below. Epigenetic inheritance is the transmission of information from an individual to its progeny without modification of the DNA sequences of genes, but rather DNA characteristics are changed, such as the DNA methylation patterns (Anway and Skinner, 2006). DNA methylation is the most studied form of epigenetics, and this methylation helps up- or downregulate several molecular events including gene transcription, X-chromosome inactivation, chromosome positioning, and repression of parasitic DNAs and imprinted genes. Imprinting, which transfers parent-specific information to progeny, involves DNA methylation of the carbon 5 position of a cytosine residue positioned next to a guanine residue, called CpG dinucleotides or islands, and is considered a major form of epigenetic modification. Imprinted genes have been hypothesized as one method for the developing animal to respond to environmental pressures (Swales and Spears, 2005). When this is considered in the context of toxicant exposures, it is proposed that chemicals can alter DNA methylation, and the consequences of this methylation are just beginning to be investigated. Epigenetic reprogramming of the genome during preimplantation may be altered due to an exogenous insult (MacPhee, 1998), resulting in permanent alteration of DNA methylation patterns (Guerrero-Bosagna et al., 2005). DNA methylation occurs at two points in development:(1) during gastrulation, which affects somatic cell development following fertilization, and (2) during gonad development, which leads to changes in methylation patterns that are sex specific and can alter the heritable epigenetic information (Anway and Skinner, 2006). As described for the mouse (Swales and Spears, 2005), mature, hypermethylated gene sequences of the parental male and female gametes converge during fertilization to form the zygote. During early embryo development, imprinted genes of the somatic and primordial germ cells (PGC) retain the parental imprints, but nonimprinted genes are demethylated. Within two weeks following fertilization, inherited parental imprinted genes of PGCs undergo demethylation; however, somatic cells maintain their imprints throughout embryo/fetal development and into adulthood. Hence, embryonic alteration of DNA methylation patterns can ultimately be expressed in adult life. Developmental changes in imprinted genes have been demonstrated in mice exposed to estrogen and other estrogenic chemicals such as DES (Newbold and Jefferson, 2006; Crews and McLachlan, 2006; McLachlan, 2001). Lactoferrin expression is regulated by estrogen in the uterine epithelial cells of adult mice, and this gene is abnormally expressed following neonatal exposure to DES. The gene continues to be abnormally expressed into adulthood, even if the animals are ovariectomized, implying that gene expression that was once estrogen dependent is now independent of estrogen regulation. The lactotransferrin gene was monitored in mice for methylation or demethylation of the regulatory elements of the gene following exposure to estrogen or DES during different periods of life (Li et al., 1997). A region upstream of the estrogen response element (ERE) of the lactoferrin gene promoter is abnormally demethylated at one CpG site in mature uteri of mice exposed to DES as neonates. The control mice, on the other hand, had normal methylation patterns. A one base pair change in the methylation pattern of the DNA can have a strong effect on gene expression, influencing which genes are transcribed or repressed, and the methylation/ demethylation patterns observed relative to the lactoferrin gene in mice exposed to DES may be linked to its ability to promote tumor formation in adult animals initially exposed as neonates (Swales and Spears, 2005; Li et al., 1997, 2003; McLachlan, 2001). Some diseases associated with imprinting errors include certain cancers and immunodeficiencies,

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centromeric region instability, facial anomalies syndrome, and Rett syndrome (Swales and Spears, 2005). Methylation pattern change can also lead to a decrease in DNA methylation. Undermethylation occurred on the 45S ribosomal RNA gene in the sperm of mice exposed to Cr (III) chloride. These changes were then associated with neoplastic and thyroid hormone (T3) changes in the offspring of exposed adults (Cheng et al., 2004). Although changes beyond the F1 generation were not investigated, these initial findings demonstrate a transgenerational effect and a chemical’s ability to potentially alter epigenetic inheritance, which could be transferred up to and beyond the F3 generation. Epigenetic transgenerational inheritance has recently been documented in the offspring of rodents exposed to EACs during gestation (Anway et al., 2005). Exposure of pregnant rats to high doses of vinclozolin (a fungicide with antiandrogenic activity) or methoxychlor (a pesticide with estrogenic, antiestrogenic, and antiandrogenic activity), during the period of embryo gonadal differentiation, resulted in decreased spermatogenic capacity and increased male infertility in the first generation of adults. These effects were transferred to the next three generations of offspring (F1–F3) as they reached adulthood in the vinclozolin treated line and through the F2 generation with the methoxychlor treated line. Moreover, the effects of vinclozolin appeared to be mediated by alterations in DNA methylation patterns through the male germlines during reprogramming. Although the doses were high, the 2005 study by Anway and colleagues provides evidence that environmental chemicals can have latent effects that are caused by changes in the epigenetic manner of inheritance in mammals. Furthermore, when the F1, F2, and F3 generations were allowed to grow and develop between 6 and 14 months of age, a greater than normal number of illnesses occurred in the mice including tumor development, prostate disease, kidney disease, testes abnormalities, immune system defects, and blood abnormalities (Anway and Skinner, 2006). The frequency of these disease phenotypes remain consistent for four generations and indicate that they are linked to the changes in methylation patterns observed in the germ cells. Several of the genes methylated in the vinclozolin exposure studies were identified and confirm an epigenetic alteration in the male germline (Chang et al., 2006). Continued work is needed to understand how these changes are associated with the transgenerational effects observed in these studies. The ultimate impacts of environmental chemicals on transgenerational effects, especially through epigenetic inheritance, are difficult to predict at this time. However, the results of the few studies described above warrant a greater need for research in this area with special consideration for hazard identification and risk assessment, especially if these effects prove to be underlying factors for common human and wildlife diseases.

18.7 CONCLUSION As discussed in this chapter, the endocrine system is integrative in nature and can be affected directly or indirectly, at many levels, in response to chemical exposures. These effects can be immediately detrimental to the organism (e.g., thyroid status during in utero development) or can be expressed later in life or in subsequent generations. Key components of the endocrine system serve to send and receive hormonal signals to and from target tissues of endocrine as well as nonendocrine organs until they are eliminated through metabolism and excretion. Among the topics chosen for inclusion in this chapter are emerging areas in endocrine research considered pivotal to assessing the full range of endocrine-related effects. Emphasis

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has also been given to different biological levels of organization ranging from molecular and cellular to whole animal, developmental, and transgenerational effects, which at the higher levels are indicative of what is happening in the real world. Due to the inherent difficulty in making causal relationships without filling in all of the links in between (i.e., altered gene expression from EACs relating to clinical disease), more research is needed in this area. We have also illustrated relevant end points and clinical signs indicative of endocrine action related to Fig. 18.1 through each of the components of the endocrine system (Table 18.1). As research findings are revealed, our understanding of how we respond to chemicals that affect the endocrine system will advance. Ultimately, such information will reveal the hazards and risks of EACs for humans and wildlife. Although the approach to risk assessment in humans focuses on the individual, and that in wildlife primarily emphasizes the population, except for endangered or charismatic species, due to the conserved nature of the endocrine system across taxa, many of the end points used to detect exposure to EACs are similar and readily translated from nonhuman animals to humans. Also, it is prudent to remember that wildlife data can be useful to estimate risk for human health. As demonstrated by this review, understanding the effects of EACs is a complicated task, because it includes myriad events from DNA regulation to population-level effects. It is our hope that this review stimulates interest in the study of consequences of exposure to EACs in humans and wildlife, encouraging scientists from different disciplines to think about the research issues facing this field and to possibly take a more global approach in this line of research.

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19 SECONDHAND SMOKE Jonathan M. Samet, Gila I. Neta and Sophia S. Wang

Extensive toxicological, experimental, and epidemiological data, largely collected since the 1950s, have established that active cigarette smoking is the major preventable cause of morbidity and mortality in the United States (U.S. Department of Health and Human Services, 1989; U.S. Department of Health and Human Services, 2004). More recently, since the 1970s, involuntary exposure to tobacco smoke has been investigated as a risk factor for disease and also found to be a cause of preventable morbidity and mortality in nonsmokers. The 1986 report of the Surgeon General on smoking and health and a report by the National Research Council (NRC), also published in 1986, comprehensively reviewed the data on involuntary exposure to tobacco smoke and reached comparable conclusions with significant public health implications (National Research Council and Committee on Passive Smoking, 1986; U.S. Department of Health and Human Services, 1986); both reports concluded that involuntary smoking causes disease in nonsmokers. Subsequently, the Environmental Protection Agency (EPA) reached a similar conclusion in its 1992 risk assessment, which classified environmental tobacco smoke as a class A carcinogen (U.S. Environmental Protection Agency, 1992). These conclusions have already had significant impact on public policy and public health. Subsequently, a now substantial body of evidence (Table 19.1) has continued to identify new diseases and other adverse effects of secondhand smoke (SHS)(California Environmental Protection Agency and Air Resources Board, 2005; California Environmental Protection Agency and Office of Environmental Health Hazard Assessment, 1997; International Agency for Research on Cancer, 2004; Scientific Committee on Tobacco and Health and HSMO, 1998; World Health Organization, 1999). The 2006 U.S. Surgeon General’s report leaves no doubt that any exposure to tobacco smoke is harmful to human health (U.S. Department of Health and Human Services, 2006). The findings on secondhand smoke and disease have been the foundation of the drive for smoke-free indoor environments and for educating parents concerning the effects of their smoking on their children’s health.


703

704 Yes/c

Yes/c

Yes/c Yes/c Yes/c

Yes/c

Yes/c Yes/c

Yes/a

Yes/a

Yes/c

Yes/c

Yes/a

Yes/c

Yes/c

Yes/c

Yes/c

Yes/c

Yes/c

Yes/c

WHO 1999

Yes/c

IARC 2004

Yes/c

Yes/c

Yes/c

Yes/c Yes/c Yes/c

Yes/c

Yes/c

Yes/c

Yes/a

Yes/c

Cal/EPAa 2005

Yes/c

Yes/c Yes/c Yes/c

Yes/c

Yes/c

Yes/c

Yes/c

Yes/c

SGR 2006

a

Only effects causally associated with SHS exposure are included.

Yes/a: association; Yes/c: cause. Table adapted from U.S. Department of Health and Human Services (2006). SGR 1984: U.S. Department of Health and Human Services (1984); SGR 1986: U.S. Department of Health and Human Services (1986); EPA 1992: U.S. Environmental Protection Agency (1992); Cal/EPA 1997: California Environmental Protection Agency and Office of Environmental Health Hazard Assessment (1997; UK 1998: Scientific Committee on Tobacco and Health and HSMO (1998); WHO 1999: World Health Organization (1999); IARC 2004: International Agency for Research on Cancer (2004); Cal/EPA 2005: California Environmental Protection Agency and Air Resources Board (2005); SGR 2006: U.S. Department of Health and Human Services (2006).

Yes/c

Yes/a Yes/c

Yes/c

Yes/a

Yes/a

Yes/a

Yes/a

Yes/c

Yes/a

Yes/c

Yes/a

Yes/c

Yes/a

UK 1998

Yes/a

Cal/EPA 1997

Increased prevalence of chronic respiratory symptoms Decrement in pulmonary function Increased occurrence of acute respiratory illnesses Increased occurrence of middle ear disease Increased severity of asthma episodes and symptoms Risk factor for new asthma Risk factor for SIDS Risk factor for lung cancer in adults Risk factor for breast cancer for younger, primarily postmenopausal women Risk factor for nasal sinus cancer Risk factor for heart disease in adults

EPA 1992

SGR 1986

SGR 1984

Adverse Effects from Exposure to Tobacco Smoke

Health Effect

TABLE 19.1

EXPOSURE TO SECONDHAND SMOKE

705

This chapter provides an overview of the evidence on secondhand smoke and its impact on the health of children and adults. It covers the conclusions of the major recent reports that have systematically evaluated the evidence. The chapter describes the findings of some key representative studies, but it is not systematic in approach given the current scope of the literature. Complete reviews are provided by the 2005 Cal/EPA and 2006 Surgeon General’s reports (California Environmental Protection Agency and Air Resources Board, 2005; U.S. Department of Health and Human Services, 2006). We also note that there has been a long-standing campaign by the tobacco industry to discredit the evidence on secondhand smoke and health, in order to maintain an apparent controversy as a basis for slowing tobacco control. These tactics are well described through research based on the industry’s own documents, obtained as a result of litigation (Tobacco Documents Online, 2006). The influence of this campaign has extended to the peer-reviewed literature, including reports on methodologic issues, exposures, epidemiological studies, risk estimates, and control measures. 19.1 EXPOSURE TO SECONDHAND SMOKE 19.1.1

Characteristics of SHS

Nonsmokers inhale SHS, the combination of the sidestream smoke that is released from the burning end of the cigarette and the mainstream smoke exhaled by the active smoker (U.S. Department of Health and Human Services, 2006). This mixture has also been referred to as environmental tobacco smoke (ETS) or SHS. The inhalation of SHS is generally referred to as passive smoking or involuntary smoking. The exposures of involuntary and active smoking differ quantitatively and, to some extent, qualitatively (International Agency for Research on Cancer, 2004). Because of the lower temperature in the burning cone of the smoldering cigarette, most partial pyrolysis products are enriched in sidestream compared to mainstream smoke. Consequently, sidestream smoke has higher concentrations of some toxic and carcinogenic substances than mainstream smoke; however, dilution by room air markedly reduces the concentrations inhaled by the involuntary smoker in comparison to those inhaled by the active smoker. Nevertheless, involuntary smoking is accompanied by exposure to toxic agents generated by tobacco combustion (International Agency for Research on Cancer, 2004; U.S. Department of Health and Human Services, 1986). 19.1.2

Secondhand Smoke Concentrations

Tobacco smoke is a complex mixture of gases and particles that contains myriad chemical species (Guerin et al., 1992; International Agency for Research on Cancer, 2004; U.S. Department of Health and Human Services, 1984). Not surprisingly, tobacco smoking in indoor environments increases levels of respirable particles, nicotine, polycyclic aromatic hydrocarbons, carbon monoxide CO, acrolein, nitrogen dioxide (NO2), and many other substances. The extent of the increase in concentrations of tobacco smoke components varies with the number of smokers, the intensity of smoking, the rate of exchange between the indoor air space and the outdoor air, and the use of air-cleaning devices. Ott (1999) has used mass balance models to characterize factors influencing concentrations of tobacco smoke indoors. Using information on the source strength (i.e., the generation of emissions by cigarettes) and on the air exchange rate, researchers can apply mass balance models to

706

SECONDHAND SMOKE

predict tobacco smoke concentrations. Such models can be used to estimate exposures and to project the consequences of control measures. Several components of cigarette smoke have been measured in indoor environments as markers of the contribution of tobacco combustion to indoor air pollution. Particles have been measured most often because both sidestream and mainstream smoke contain high concentrations of particles in the respirable size range (International Agency for Research on Cancer, 2004; U.S. Department of Health and Human Services, 1986; 2006). Particles are a nonspecific marker of tobacco smoke contamination, however, because numerous sources other than tobacco combustion add particles to indoor air. Other more specific markers have also been measured, including nicotine, solanesol, and ultraviolet light (UV) absorption of particulate matter (Guerin et al., 1992). Nicotine, which is present in the gas phase in secondhand smoke, is usually measured with passive diffusion badges (Guerin et al., 1992; Leaderer and Hammond, 1991; U.S. Department of Health and Human Services, 2006). Nicotine has become the principal marker because of its specificity to cigarette smoke and the ease of measurement with the passive monitors. Many studies of levels of SHS components have been conducted in public buildings; fewer studies have been conducted in homes and offices (U.S. Department of Health and Human Services, 1986, 2006). The contribution of various environments to personal exposure to tobacco smoke varies with the time–activity pattern, namely, the distribution of time spent in different locations. Time–activity patterns may heavily influence lung airway exposures in particular environments for certain groups of individuals. For example, exposure in the home predominates for infants who do not attend day care (Harlos et al., 1987). For adults residing with nonsmokers, the workplace may be the principal location where exposure takes place. A nationwide study assessed exposures of nonsmokers in 16 metropolitan areas of the United States (Jenkins et al., 1996). This study, involving 100 persons in each location, was directed at workplace exposure and included measurements of respirable particulate matter and other markers. The results showed that in 1993 and 1994, exposures to SHS in the home were generally much greater than those in the workplace. The contribution of smoking in the home to indoor air pollution has been demonstrated by studies using personal monitoring and monitoring of homes for respirable particles. In one of the earliest studies, Spengler and Tosteson (1981) monitored homes in six U.S. cities for respirable particle concentrations over several years and found that a smoker of one pack of cigarettes daily contributed about 20 mg/m3 to 24 h indoor particle concentrations. In homes with two or more heavy smokers, this study showed that the pre-1987 24 h National Ambient Air Quality Standard (NAAQS) of 260 mg/m3 for total suspended particles could be exceeded. Because cigarettes are not smoked uniformly over the day, higher peak concentrations must occur when cigarettes are actually smoked. Spengler et al. (1985) measured the personal exposures to respirable particles sustained by nonsmoking adults in two rural Tennessee communities. The mean 24 h exposures were substantially higher for those exposed to smoke at home: 64 mg/m3 for those exposed versus 36 mg/m3 for those not exposed. Nicotine levels have now been measured in multiple homes in the United States, as shown in Fig. 19.1 (Emmons et al., 2001; Hammond et al., 1989; Henderson et al., 1989; Jenkins et al., 1996; Leaderer and Hammond, 1991; Marbury et al., 1993; U.S. Department of Health and Human Services, 2006). In homes with smokers, mean values range about 2–5 mg/m3 in the various studies, but maximum values in some homes are much higher. Additionally, these measures reflect the average concentration across the time of measurement, but not the values when nonsmokers are actually being exposed.


707

FIGURE 19.1 Concentration of nicotine in homes of U.S. smokers (U.S. Department of Health and Human Services, 2006).

The Total Exposure Assessment Methodology (TEAM) study, conducted by the U.S. Environmental Protection Agency, provided extensive data on concentrations of 20 volatile organic compounds in a sample of homes in several communities (Wallace and Pellizzari, 1987). Cigarette combustion is a strong source of many volatile organic compounds. Indoor monitoring showed increased concentrations of benzene, xylenes, ethylbenzene, and styrene in homes with smokers compared to homes without smokers. Figures 19.2 and 19.3 illustrate that more extensive information is now available on levels of SHS components in public buildings and workplaces of various types (U.S. Department of Health and Human Services, 2006). Monitoring in locations where smoking may be intense, such as bars and restaurants, has generally shown substantial elevations of particles and other markers of smoke pollution while smoking is taking place. For example, Repace and Lowrey (1980) in an early study used a portable piezobalance to sample aerosols in restaurants, bars, and other locations. In the places sampled, respirable particulate levels ranged up to 700 mg/m3, and the levels varied with the intensity of smoking (California Environmental Protection Agency and Office of Environmental Health Hazard Assessment, 1997; Guerin et al., 1992; National Research Council and Committee on Passive Smoking, 1986). Studies summarized in the 2006 report of the Surgeon General (U.S. Department of Health and Human Services, 2006) show the widespread presence of nicotine in workplaces and other locations with smoking allowed (Figs 19.2 and 19.3) and the potential for maximum concentrations to be extremely high. Monitoring studies document the effectiveness of workplace smoking policies for sharply reducing nicotine concentrations. Recent studies indicate low concentrations in many workplace settings, reflecting declining smoking prevalence in recent years and changing practices of smoking in the workplace. Using passive nicotine samplers, Hammond (Hammond, 1999) showed that worksite smoking policies can sharply reduce SHS exposure. Transportation environments may also be polluted by cigarette smoking. Contamination of air in trains, buses, automobiles, airplanes, and submarines has been documented (National Research Council and Committee on Passive Smoking, 1986; U.S. Department of Health and Human Services, 1989). A NRC report (National Research Council and

FIGURE 19.2 Occupational exposures to nicotine among groups of nonsmoking office workers (U.S. Department of Health and Human Services, 2006).

FIGURE 19.3 Average concentrations of nicotine in homes, offices, other workplaces, and restaurants where smoking is permitted (U.S. Department of Health and Human Services, 2006). 708


709

Committee on Airliner Cabin Environment Safety Committee, 1986) on air quality in airliners summarized studies for tobacco smoke pollutants in commercial aircraft. In one study, during a single flight, the NO2 concentration varied with the number of passengers with a lighted cigarette. In another study, respirable particles in the smoking section were measured at concentrations five or more times higher than in the nonsmoking section. Peaks as high as 1000 mg/m3 were measured in the smoking section. (Mattson et al., 1989) used personal exposure monitors to assess nicotine exposures of passengers and flight attendants. All persons were exposed to nicotine, even if seated in the nonsmoking portion of the cabin. These studies are now of historical interest only as almost all commercial flights worldwide are smoke free. Automobiles, however, are potential sites of high levels of exposure: preliminary data from a study in Greece (Vardavas et al., 2006) show that levels of concentrations can reach excessive heights when a smoker in a car exposes others to secondhand smoke. 19.1.3

Biological Markers of Exposure

Biological markers can be used to describe the prevalence of exposure to secondhand smoke, to investigate the dosimetry of involuntary smoking, and to validate questionnaire-based measures of exposure. In both active and involuntary smokers, the detection of tobacco smoke components or their metabolites in body fluids or alveolar air provides evidence of exposure to tobacco smoke, and levels of these markers can be used to gauge the intensity of exposure. The risk of involuntary smoking has also been estimated by comparing levels of biological markers in active and involuntary smokers. At present, the most sensitive and specific markers for tobacco smoke exposure are nicotine and its metabolite, cotinine (International Agency for Research on Cancer, 2004; Jarvis and Russell, 1984; U.S. Department of Health and Human Services, 1988, 2006). Neither nicotine nor cotinine is usually present in body fluids in the absence of exposure to tobacco smoke, although unusually large intakes of some foods could produce measurable levels of nicotine and cotinine (Idle, 1990). Cotinine, formed by oxidation of nicotine by cytochrome P450, is one of the several primary metabolites of nicotine (U.S. Department of Health and Human Services, 1988). Cotinine itself is extensively metabolized, and only about 17% of cotinine is excreted unchanged in the urine (International Agency for Research on Cancer, 2004). Because the circulating half-life of nicotine is generally shorter than 2 h (Rosenberg et al., 1980), nicotine concentrations in body fluids reflect more recent exposures. Nicotine can be measured in hair, as it is incorporated into the growing hair. By using several centimeters of hair, the level of nicotine reflects exposure over several weeks (Jaakkola and Jaakkola, 1997). In contrast to the short half-life of nicotine in the blood, cotinine has a half-life of about 10 h in the blood or plasma of active smokers (U.S. Department of Health and Human Services, 2006) and of about 20 h in nonsmokers (Kyerematen et al., 1982; U.S. Department of Health and Human Services, 2006), and hence, cotinine levels in blood, urine, or saliva provide information about exposure to tobacco smoke of involuntary smokers over periods of several days (Turner et al., 1987; Wall et al., 1988). Concerns about nonspecificity of cotinine, arising from eating nicotine-containing foods, have been set aside (Benowitz, 1996). Thiocyanate concentration in body fluids, concentration of CO in expired air, and carboxyhemoglobin level distinguish active smokers from nonsmokers but are not as sensitive and specific as cotinine for assessing involuntary exposure to tobacco smoke (Jarvis and Russell, 1984; U.S. Department of Health and Human Services, 2006).

710

SECONDHAND SMOKE

FIGURE 19.4 Serum cotinine geometric means (95% CI) for U.S. nonsmokers by study interval: exposure of nonsmokers in the U.S. population to SHS, 1988–2002. The data are plotted at the approximate midpoint for four separate time intervals: 1988–1991 (NHANES III, phase (1), 1991–1994 (NHANES III, phase (2), 1999–2000, and 2001–2002 (Pirkle et al., 2006).

Cotinine levels have been measured in adult nonsmokers and in children (U.S. Department of Health and Human Services, 2006). In the studies of adult nonsmokers, exposures at home, in the workplace, and in other settings determined cotinine concentrations in urine and saliva. The cotinine levels in involuntary smokers ranged from less than 1% to about 8% of cotinine levels measured in active smokers. Smoking by parents is the predominant determinant of the cotinine levels in their children. In 1988–1991, using liquid chromatography–mass spectrometry as the assay method, it was found that 88% of nonsmokers had a detectable level of serum cotinine (Pirkle et al., 1996, 2006; U.S. Department of Health and Human Services, 2006). Cotinine levels in this national sample increased with the number of smokers in the household and the hours exposed in the workplace. In subsequent phases of NHANES, the proportions of participants with a detectable level of cotinine and the mean level have dropped substantially (Fig. 19.4). The results of studies on biological markers have important implications for research on involuntary smoking and add to the biological plausibility of associations between involuntary smoking and diseases documented in epidemiological studies (Benowitz, 1996). The data on marker levels provide ample evidence that involuntary exposure leads to absorption, circulation, and excretion of tobacco smoke components. The studies of biological markers also confirm the high prevalence of involuntary smoking, as ascertained by questionnaire (Benowitz, 1996; Coultas et al., 1987; Pirkle et al., 1996). The observed correlations between reported exposures and levels of markers suggest that questionnaire methods for assessing recent exposure have some validity. 19.1.4

Exposure Assessment

The information on the health effects of involuntary smoking has been largely derived from observational epidemiological studies. In these studies, exposure to SHS has been estimated primarily by responses to questionnaires concerning the smoking habits of household members or fellow employees; attempts have been made to quantitate exposure by determining the number of cigarettes smoked by family members and the duration of

HEALTH EFFECTS OF INVOLUNTARY SMOKING IN CHILDREN

711

exposure. Biomarkers have also been used in some studies. Limitations of the questionnaire approach were discussed extensively in the 1986 report of the Surgeon General and again in the 2006 report (U.S. Department of Health and Human Services, 1986, 2006). The potential for information bias to introduce positive associations of SHS exposure with disease risk has been a focus of debate. A number of studies have addressed characteristics of questionnaires and biological markers for assessing exposure to SHS. Two studies evaluated the reliability of questionnaires on lifetime exposure (Coultas et al., 1989; Lubin, 1999; Pron et al., 1988). Both showed a high degree of repeatability for questions concerning whether a spouse had smoked but a lower reliability for responses concerning quantitative aspects of exposure. Several studies have assessed the validity of subjects’ reports on smoking by parents and spouses. Sandler and Shore (1986) compared responses on parents’ smoking given by cases and controls with responses given by the parents or siblings of the index subjects. Concordance was high for whether the parents had ever smoked. Responses concerning numbers of cigarettes smoked did not agree as highly. In a follow-up study of a nationwide sample, children’s responses on smoking by their deceased parents closely agreed with the information given 10 years previously by the parents (McLaughlin et al., 1987). A number of studies have shown that people correctly report the smoking habits of their spouses (U.S. Department of Health and Human Services, 1990b). In a study of nonsmokers in Buffalo, index subjects’ reports agreed well with reports from parents or siblings, spouse or children, and coworkers concerning exposure during childhood, at home, and at work, respectively (Cummings et al., 1989). Coghlin et al. (1989) used a passive nicotine monitor as well as a questionnaire and diary approach for characterizing exposure to SHS. In a sample of 19 volunteers, they found a strong correlation between the monitored nicotine exposure and a questionnaire-based index; the sampling lasted only a week, however, and the diary method would be too cumbersome to implement among all participants in a large epidemiological study. Although biological markers have provided important evidence of population exposures, the utility of cotinine as an indicator of individual exposure has been questioned. Idle (1990) has reviewed the complex metabolism of nicotine and the many factors affecting the relationship between exposure to atmospheric nicotine and the concentration of cotinine in body fluids. He cautions against using any single determination of cotinine as a measure of exposure. Several epidemiological studies support this concern about the limited validity of a single measurement of cotinine. Spot cotinine levels are not tightly predicted by questionnaire measures of exposures (Coultas et al., 1989; Cummings et al., 1990), and cotinine levels are highly variable at any particular level of smoking in a household (Coultas et al., 1990). Thus, questionnaires remain the best method for characterizing usual exposure to SHS. However, biological markers and personal monitoring offer complementary approaches for developing more accurate exposure estimates for quantifying dose and judging the extent of misclassification introduced by questionnaires. 19.2 HEALTH EFFECTS OF INVOLUNTARY SMOKING IN CHILDREN 19.2.1

Fetal Effects

Researchers have demonstrated that active smoking by mothers results in a variety of adverse health effects in children. Some of the health effects predominantly result from transplacental exposure of the fetus to tobacco smoke components. Recently, studies have also

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investigated and demonstrated associations between adverse health effects in children and exposure to SHS. For example, paternal smoking in the presence of a pregnant mother may lead to perinatal health effects manifested upon birth of the baby, and either maternal or paternal smoking in the presence of a newborn child may lead to postnatal health effects in the developing child. Health effects on the fetus resulting from SHS include fetal growth effects (decreased birth weight, growth retardation, or prematurity), fetal loss (spontaneous abortion and perinatal mortality), and congenital malformations. Health effects on the child postnatally, resulting from SHS exposure either to the fetus or to the newborn child, include sudden infant death syndrome (SIDS) and adverse effects on neuropsychological development and physical growth. Possible longer term health effects of fetal SHS exposure include childhood cancers of the brain, leukemia, and lymphomas, among others. 19.2.2

Biological Plausibility

This topic receives extensive coverage in the 2006 Surgeon General’s report (U.S. Department of Health and Human Services, 2006). Secondhand smoke could plausibly have adverse health effects at one or more steps of the developmental and reproductive processes, particularly during critical periods of susceptibility. Presumably, particular SHS components reach target sites and have toxic effects. For example, fetal exposure to CO and nicotine due to SHS may increase risk for perinatal health effects. CO in SHS may contribute to increased concentrations of CO and carboxyhemoglobin in the fetus, and the fetus may not be able to physiologically compensate for the reduced oxygen delivery (U.S. Department of Health and Human Services, 1980), leading to fetal hypoxia. With regard to SIDS, a number of mechanisms have been identified—arousal failure, inadequate cardiorespiratory compensatory motor responses, and sleep apnea—attributable to developmental abnormalities in the brainstem and autonomic nervous system (U.S. Department of Health and Human Services, 2006). Animal models indicate lasting effects on brain nicotine receptors (Slotkin, 2004). Animal models also indicate neural cellular effects from postnatal exposure that could underlie the link between paternal smoking and increased risk for SIDS (U.S. Department of Health and Human Services, 2006). Association between SHS and childhood cancers is biologically plausible due to the presence of carcinogenic tobacco smoke components or metabolites, such as benzene, nitrosamines, urethane, and radioactive compounds, at organ sites of the cancers. In animal studies, neurogenic tumors as well as other tumors were induced after transplacental exposure to a number of compounds present in tobacco smoke, including several nitrosamines. Smoking metabolites such as thiocyanate have also been found in fetal blood (Bottoms et al., 1982; Coghlin et al., 1991) and amniotic fluid (Andersen et al., 1982; Smith et al., 1982) of nonsmoking women exposed to tobacco smoke. Moreover, Huel et al. (1989) measured aryl hydrocarbon hydroxylase activity in human placenta of involuntary tobacco smokers; levels were increased in placentas of women passively exposed to tobacco smoke. 19.2.3

Nonfatal Perinatal Health Effects

19.2.3.1 Fetal Growth Most studies have used paternal smoking as the exposure measure to assess the association between SHS exposure and nonfatal perinatal health effects, such as reduced fetal growth. Low birth weight was first reported in 1957 to be


713

associated with maternal smoking during pregnancy (U.S. Department of Health and Human Services, 1980). On average, birth weight is reduced by about 250 g for newborns whose mothers smoked during pregnancy (U.S. Department of Health and Human Services, 2001; U.S. Department of Health and Human Services, 2004). Extensive studies have since been conducted to assess SHS exposure and birth weight, with consideration of gestational age at delivery, multiple births, maternal age, race, parity, maternal smoking, socioeconomic status, pregnancy history (U.S. Department of Health and Human Services, 2001, 2006). Exposures have been measured with questionnaires that assess home and work exposure, and in some studies, with the use of biomarkers. Studies generally find lower birth weight for infants of nonsmoking women passively exposed to tobacco smoke during pregnancy (Rubin et al., 1986; U.S. Department of Health and Human Services, 2006). For example, Haddow et al. (1988) used cotinine as a biomarker to measure exposure to SHS; they also adequately controlled for potential confounders. SHS exposure was defined as cotinine levels of 1.1–9.9 ng/mL in the fetus born to a nonsmoking mother. Their study demonstrated a decrease of 100 g in birth weight for fetuses exposed to SHS. Other biomarker studies (Eskenazi and Bergmann, 1995; Eskenazi and Trupin, 1995; Martinez et al., 1994) support the findings of Haddow et al. (1988). Other epidemiologic studies assessed SHS exposure from multiple sources through questionnaire (Mainous and Hueston, 1994; Rebagliato et al., 1995; Roquer et al., 1995). While not using a method as specific or sensitive as cotinine measurements, these studies still demonstrated decreases in mean birth weights after adjustment for confounders (20–40 g). In a 1999 meta-analysis, Windham et al. (1999) found a mean reduction of about 3 g for infants of nonsmoking mothers exposed to SHS during pregnancy. There was also a slightly increased risk for low birth weight. The 2006 report of the U.S. Surgeon General (U.S. Department of Health and Human Services, 2006) concluded that these effects of SHS were causal. 19.2.3.2 Other Effects Other nonfatal perinatal health effects possibly associated with SHS are growth retardation and congenital malformations. Martin and Bracken (1986) demonstrated a strong association with growth retardation in their 1986 study. Later studies (Mainous and Hueston, 1994; Roquer et al., 1995) supported this finding; however, these studies had small sample sizes and did not control for potential confounders. A few studies (Savitz et al., 1991; Seidman et al., 1990; Zhang et al., 1992) have been conducted to assess the association between paternal smoking and congenital malformations. The most consistent associations appear with the central nervous system or neural tube defects. However, due to possible effects of active smoke on the sperm, a causal association between SHS and congenital malformations cannot be concluded. 19.2.4

Fetal Perinatal Health Effects

SHS exposure to the fetus during its development may lead to fatal perinatal health effects such as spontaneous abortion and perinatal mortality. Very few studies have examined the association between SHS exposure and perinatal death. Eight studies have examined neonatal mortality in relation to paternal smoking, and a few supported an increase in risk (Ahlborg and Bodin, 1991; Comstock and Lundin, 1967; Lindbohm et al., 1991; Mau and Netter, 1974).

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19.2.5

SECONDHAND SMOKE

Postnatal Health Effects

SHS exposure due to maternal or paternal smoking may lead to postnatal health effects related to SIDS, physical development, decrements in cognition and behavior, and cancers. 19.2.5.1 SIDS Sudden infant death syndrome (SIDS) is the sudden, unexplained death of an infant under 1 year of age. SIDS has been causally associated with maternal smoking during pregnancy (U.S. Department of Health and Human Services, 2001, 2004) and now with SHS exposure after birth (California Environmental Protection Agency and Air Resources Board, 2005; U.S. Department of Health and Human Services, 2006). The 2006 Surgeon General’s report cites 13 studies directed at the association between SIDS and postnatal SHS exposure. Ten studies have addressed the association between postpartum maternal smoking and SIDS, and nine studies considered paternal smoking. While maternal smoking during pregnancy has been causally associated with SIDS, these studies measured maternal smoking after pregnancy, along with paternal smoking and household smoking generally. Effects of SHS exposure after birth and maternal smoking during pregnancy cannot be readily separated in many of these studies, but paternal smoking involves SHS exposure with the potential to avoid the complicating consequences of smoking during pregnancy. Previously, maternal smoking during pregnancy had been causally linked to SIDS (U.S. Department of Health and Human Services, 2001), but separating the effects of prenatal exposure from maternal smoking during pregnancy and of postnatal exposure had been difficult. The evidence reviewed in the 2006 Surgeon General’s report (U.S. Department of Health and Human Services, 2006) showed that postnatal maternal smoking and paternal smoking were associated with increased risk of SIDS. Risks were increased by 50% to more than 100%, and several studies showed a dose–response relationship with level of SHS exposure. One study found increased risk for smoking by adults in the same room with the child (Klonoff-Cohen et al., 1995). Both the 2005 Cal/EPA and the 2006 Surgeon General’s reports (California Environmental Protection Agency and Air Resources Board, 2005; U.S. Department of Health and Human Services, 2006) concluded that a causal relationship exists between SHS exposure and SIDS. In reaching their conclusion, the reports noted not only the epidemiological evidence but also the findings of animal models that indicate potential mechanisms. The Cal/EPA report estimates that 10% of SIDS deaths are attributable to SHS exposure. 19.2.5.2 Cognition and Behavior While it is biologically plausible that SHS affects a child’s neuropsychological development, perhaps through nicotine’s effect on the central nervous system and through the effect of chronic exposure to CO, research on this potential consequence of SHS exposure needs to examine this relationship independent of prenatal exposure and maternal active smoking. Research on this topic also needs to carefully take into account potential confounding factors from the correlates of SHS exposure. Furthermore, cognition and behavior are also measured through a variety of tests, making direct comparisons between studies difficult. The 2006 Surgeon General’s report (U.S. Department of Health and Human Services, 2006) cited 12 epidemiologic studies that addressed SHS exposure and cognitive development in children. Of these, eight found associations between measures of involuntary smoking and children’s cognitive development. However, the evidence was heterogeneous and consistency could not be assessed. The evidence was judged to be inadequate to infer the presence or absence of a causal relationship.


19.2.6

715

Childhood Cancers

A causal association between involuntary smoking and childhood cancer derives biological plausibility from evidence of transplacental carcinogenesis in animal studies and in humans. Notably, cancers develop in offspring of rabbits and monkeys when the oncogenic compound ethylnitrosourea is administered to pregnant mothers (Rice et al., 1989; Stavrou et al., 1984), and in humans, transplacental carcinogenesis is well established between diethylstilbesterol exposure in pregnant mothers and development of vaginal clear cell adenocarcinoma among their daughters (Vessey, 1989). Evidence for transplacental carcinogenesis with involuntary smoking includes increased detection levels of smoking metabolites such as thiocyanate in fetal blood (Bottoms et al., 1982; Coghlin et al., 1991) and amniotic fluid (Andresen et al., 1982; Smith et al., 1982) compared to unexposed nonsmoking women. Moreover, genotoxic effects as measured by deletions within the housekeeping gene, hypoxanthine guanine phosphoribosyl transferase (HPRT), have been documented in cord blood of newborns from exposed mothers (Finette et al., 1998; International Agency for Research on Cancer, 1986). Most epidemiologic studies that have evaluated the association between involuntary smoking to the mother and childhood cancer among their offspring are case-control studies that have measured exposure based on the smoking habits of the father; few have included relevant exposures outside the home. Exposure assessment via the father, however, cannot distinguish the resulting cancers as being caused by involuntary smoking exposure in the mother or due to active smoking in the father that has resulted in DNA damage to the father’s sperm (U.S. Department of Health and Human Services, 2006). Further, distinguishing the effects of prenatal and postnatal secondhand smoke is difficult due to their high correlation. Finally, not all studies excluded mothers who were active smokers or accounted for other cancer risk factors such as maternal X-rays, sodium nitrite consumption, and drug use. It should be noted, however, that active maternal smoking during pregnancy has not been established as a causal risk factor for childhood cancer (Boffetta et al., 2000). Early evidence for an association between involuntary smoking in the mother and childhood cancer came from a 1982 report for childhood brain cancer (Preston-Martin et al., 1982). Results from studies published since then, particularly for all cancers, have provided mixed results. One meta-analysis by Sorahan et al. (1997b) reported a 1.2-fold increased risk for all cancers with paternal smoking, but recent reviews have not consistently supported this association (International Agency for Research on Cancer, 2004; Sasco and Vainio, 1999; Tredaniel et al., 1994). A cohort study also did not support an association between paternal smoking and childhood cancer (Seersholm et al., 1997). The evidence for specific cancers, notably leukemia, lymphoma, and central nervous system tumors, are more limited but some reveal more consistent patterns than the evidence for all cancers combined. 19.2.6.1 Brain Tumors Brain tumors are the most extensively studied childhood cancer in relation to involuntary maternal smoking. The association between involuntary maternal smoking and brain tumors is biologically plausible due to endogenously formed N-nitroso precursors found in SHS. Associations between paternal smoking and risk of brain tumors have been demonstrated in a number of studies (Preston-Martin et al., 1982; Filippini et al., 1994; McCredie et al., 1994; Sorahan et al., 1997a, 1997b) with statistically significant odds ratios ranging from 1.5 to 2.2 in some but not in others (Bunin et al., 1994; Gold et al., 1993; Howe et al., 1989; Kuijten et al., 1990; Norman et al., 1996a, 1996b).

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19.2.6.2 Leukemia The hypothesized pathway of transplacental carcinogenesis for leukemia resides in the leukemogen benzene, a component of secondhand smoke. Casecontrol studies evaluating the association between childhood leukemia and paternal smoking, however, have not been consistent. Most studies do not support an association (Brondum et al., 1999; Infante-Rivard et al., 2000; John et al., 1991; Magnani et al., 1990; Shu et al., 1996) but two studies (Ji et al., 1997; Sorahan et al., 2001), which distinguished between acute lymphocytic leukemia (ALL) and non-ALL, found increased risks for ALL at the highest levels of paternal smoking prior to conception. Although the study by Sorahan et al. (2001) included both smoking and nonsmoking mothers, results by Ji et al. (1997) were derived from nonsmokers. 19.2.6.3 Lymphomas Although it is suggestive, the evidence for involuntary smoking and childhood lymphoma remains sparse. Ji et al. (1997) present the strongest evidence to date for paternal smoking and childhood lymphoma among nonsmokers. Although Sorahan et al. (2001) and Magnani et al. (1990) also demonstrate increased risks, their studies do not delineate between smoking and nonsmoking mothers. Two studies also report no association between involuntary smoking and childhood lymphoma (Sorahan et al., 1995, 1997b). In summary, the data to date suggest an association between secondhand smoke exposure and childhood cancer and in particular for childhood brain tumors, leukemias, and lymphomas. However, the evidence is not yet sufficient to establish a causal relationship, and the effects of exposures during pregnancy and during infancy cannot be separated at the present time. 19.2.7

Lower Respiratory Tract Illnesses in Childhood

Studies of involuntary smoking, particularly maternal smoking, and lower respiratory illnesses in childhood, including bronchitis and pneumonia, provided some of the earliest evidence on adverse effects of SHS (Colley et al., 1974a, 1974b; Harlap and Davies, 1974). Presumably this association represents an increase in frequency or severity of illnesses that are infectious in etiology and not a direct response of the lung to toxic components of SHS. Investigations conducted throughout the world have demonstrated an increased risk of lower respiratory tract illness in infants with smoking parents (California Environmental Protection Agency and Air Resources Board, 2005; U.S. Department of Health and Human Services, 2006). These studies indicate a significantly increased frequency of bronchitis and pneumonia during the first year of life of children with smoking parents. The 2006 Surgeon General’s report (U.S. Department of Health and Human Services, 2006) includes a quantitative review of this information, combining data from over 50 studies. Overall, there is an approximate 50% increase in illness risk if either parent smoked, with the risk for maternal smoking being somewhat higher (Fig. 19.5). Although the health outcome measures have varied somewhat among the studies, the relative risks associated with involuntary smoking were similar, and dose–response relationships with extent of parental smoking were demonstrable. Although most of the studies have shown that maternal smoking, rather than paternal smoking, underlies the increased risk of lower respiratory tract illnesses, studies from China and elsewhere show that paternal smoking alone can increase incidence of lower respiratory illness (Strachan and Cook, 1997; U.S. Department of Health and Human Services, 2006). In these studies, an effect of passive smoking has not been readily identified after the first year of life. During the first year of life, the strength of its effect may reflect higher exposures consequent to the time–activity patterns of young infants, which place them in close proximity to cigarettes smoked by their mothers.


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FIGURE 19.5 Odds ratios for the effect of smoking by either parent on lower respiratory illnesses during infancy (U.S. Department of Health and Human Services, 2006).

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19.2.8

Respiratory Symptoms and Illness in Children

The evidence on respiratory symptoms and illnesses in children and SHS exposure is now coming from numerous surveys and also from cohort studies. Data from surveys, largely of school children, demonstrate a greater frequency of the most common respiratory symptoms, cough, phlegm, and wheeze, in the children of smokers (California Environmental Protection Agency and Air Resources Board, 2005; California Environmental Protection Agency and Office of Environmental Health Hazard Assessment, 1997; Cook and Strachan, 1997a; U.S. Department of Health and Human Services, 1986, 2006) (Table 19.2). In these studies, the subjects have generally been school children, and the effects of parental smoking TABLE 19.2 Summary of Pooled Random Effects (Odds Ratios) Associated with Parental Smoking Restricted to Studies of Children Aged 11 Years (U.S. Department of Health and Human Services, 2006) Odds Ratio for Smoking (95% CI) Symptom Asthma

Numbers of Studies

Either Parent

13

1.18 (1.06–1.31)

One Parent

Both Parents

Mother Only

Father Only

Insufficient studies 5

1.47 (1.29–1.68)

7

1.31 (1.15–1.50)

4 Wheezea

15

1.13 (0.99–1.29) 1.27 (1.16–1.38)

4

1.21 (1.10–1.45)

5

1.41 (1.16–1.71)

8

1.26 (1.15–1.36)

5 Cough

13 4 5 4

1.10 (1.02–1.20) 1.28 (1.13–1.44) 1.17 (0.84–1.61) 1.85 (1.29–2.64) 1.07 (0.91–1.24)

3

1.12 (0.95–1.38)

Note: The symptoms “phlegm” and “breathlessness” were included in this table because of an insufficient number of studies. a

Excluded the European Communities Study, which had a pooled odds ratio of 1.20.


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have been examined. Thus, the less prominent effects of passive smoking, in comparison to the studies of lower respiratory illness in infants, may reflect lower exposures to SHS by older children who spend less time with their parents. By the mid-1980s, results from several large studies provided convincing evidence that involuntary exposure to SHS increases the occurrence of cough and phlegm in the children of smokers, although earlier data from smaller studies had been ambiguous. For example, in a study of 10,000 school children in six U.S. communities, smoking by parents increased the frequency of persistent cough in their children by about 30% (Ware et al., 1984). The effect of parental smoking was derived primarily from smoking by the mother. Charlton (1984) conducted a survey on cigarette smoking that included 15,709 English children aged 8–19 years. In the nonsmoking children, the prevalence of frequent cough was significantly higher if either the father or the mother smoked. For the symptom of chronic wheeze, the preponderance of the early evidence also indicated an excess associated with involuntary smoking.Inasurveyof650schoolchildreninBoston,oneofthefirststudiesonthisassociation, persistent wheezing was the most frequent symptom (Weiss et al., 1980); the prevalence of persistent wheezing increased significantly as the number of smoking parents increased. In the SixCitiesStudyofchildren,theprevalenceofpersistentwheezingduringthepreviousyearwas significantly increased if the mother smoked (Ware et al., 1984). 19.2.9

Childhood Asthma

Although involuntary exposure to tobacco smoke has been associated with the symptom of wheeze, evidence for association of involuntary smoking with childhood asthma was initially conflicting. Exposure to SHS might cause asthma as a long-term consequence of the increased occurrence of lower respiratory infection in early childhood or through other pathophysiological mechanisms including inflammation of the respiratory epithelium (Tager et al., 1988; U.S. Department of Health and Human Services, 2006). The effect of SHS may also reflect, in part, the consequences of in utero exposure. Assessment of airways responsiveness shortly after birth has shown that infants whose mothers smoke during pregnancy have increased airways responsiveness compared with those whose mothers do not smoke (U.S. Department of Health and Human Services, 2006). Maternal smoking during pregnancy also reduced ventilatory function measured shortly after birth (Hanrahan et al., 1992). These observations suggest that in utero exposures from maternal smoking may affect lung development, perhaps reducing relative airways size. Additionally, childhood asthma is considered to have a strong genetic basis, and SHS exposure may act to increase or hasten incidence in a genetically predisposed subgroup of the population. While the underlying mechanisms remain to be fully characterized, the epidemiologic evidence linking SHS exposure and childhood asthma is substantial (California Environmental Protection Agency and Air Resources Board, 2005; Cook and Strachan, 1997a, 1997b; U.S. Department of Health and Human Services, 2006). There is evidence relevant to the causation of asthma and to the effect of SHS on the status of children with asthma. The 2006 report of the Surgeon General (U.S. Department of Health and Human Services, 2006) provides a full review of the evidence, considering the large number of cross-sectional studies and the smaller number of cohort studies. The cross-sectional studies cannot directly address SHS exposure as a cause of asthma onset because the existence of prevalent asthma reflects both incidence and maintenance of the asthmatic condition. Nonetheless, the prevalence studies provide firm evidence that prevalent asthma is associated with SHS exposure at home (Table 19.2). The 2006 report considered 41 cross-sectional studies with

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quantitative risk information. Overall, if either parent smoked, the pooled odds ratio was 1.23, compared to neither smoking. Household exposure to SHS was also associated with wheeze. The evidence was judged to be sufficient to infer a causal relationship between parental smoking and ever having asthma (U.S. Department of Health and Human Services, 2006). The report separately reviewed the seven cohort studies that addressed asthma incidence and also 21 case-control studies. Interpretation of the cohort study findings is complicated by the array of outcome measures and heterogeneity of effect by age of the children. The quantitative meta-analysis yielded a pooled odds ratio of 1.31, statistically significant, for children during the first 5–7 years of life; for the school years, the estimate was only 1.13. The case-control studies of prevalent asthma showed a 40% increase in association with smoking by either parent, a 50% increase for maternal smoking, but no increase for paternal smoking. Acknowledging the complexities of interpreting the cohort data, the 2006 report concluded that the evidence was suggestive, but not sufficient to infer a causal relationship between SHS exposure from parental smoking and onset of childhood asthma. The report also noted that SHS exposure can exacerbate childhood asthma (U.S. Department of Health and Human Services, 2006). 19.2.10

Lung Growth and Development

From gestation through adolescence, the lung goes through a complex process of maturation and growth that may be adversely affected by environmental agents, including SHS. On the basis of the primarily cross-sectional data available at the time, the 1984 report of the Surgeon General (U.S. Department of Health and Human Services, 1984) concluded that the children of smoking parents in comparison with those of nonsmokers had small reductions of lung function, but the long-term consequences of these changes were regarded as unknown. In the 2 years between the 1984 and the 1986 reports, sufficient longitudinal evidence was accumulated to support the conclusion in the 1986 report (U.S. Department of Health and Human Services, 1986) that involuntary smoking reduces the rate of lung function growth during childhood. Subsequently, substantial additional evidence has been obtained from many cross-sectional studies and some cohort studies. The evidence consistently shows that children exposed to SHS in their homes have reduced ventilatory function, as assessed by spirometry. Findings from cohort studies imply that the reduction comes from a reduced rate of lung growth. Meta-analyses of the cross-sectional data provide an indication of the magnitude of the effect of SHS exposure on lung function (Cook et al., 1998; U.S. Department of Health and Human Services, 2006). The 2006 Surgeon General’s report pooled 26 studies; the effect of SHS exposure was greatest for flow measures (the mid-expiratory and end-expiratory flow rates), approximately 4–5%, and less for the forced expiratory volume in 1 s (FEV1), approximately 1%. The effect of SHS exposure was greatest if both parents smoked and was robust to adjustment for potential confounding factors. The cohort studies show that lung function during childhood is adversely affected by maternal smoking during pregnancy and further impaired by exposure after birth. Studies of lung function shortly after birth show increased airways resistance and airways responsiveness for children exposed in utero (U.S. Department of Health and Human Services, 2006). These in utero effects appear to have implications for later lung growth and development. 19.2.11

SHS and Middle Ear Disease in Children

Otitis media (OM) is one of the most frequent diseases diagnosed in children at outpatient facilities. OM occurs as a result of dysfunction in the eustachian tube; serious OM results when


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serous fluid effuses into the middle ear, and acute OM results when the serous fluid effused into the middle ear becomes infected. All stages of OM lead to varying degrees of hearing loss. There are four biologically plausible mechanisms by which SHS could lead to middle ear disease in children. First, SHS exposure could lead to decreased mucociliary clearance, increasing possible risk of dysfunction in the eustachian tube. Second, SHS may decrease eustachian tube patency due to adenoidal hyperplasia, a known risk factor for OM. Third, SHS may also decrease patency as a result of SHS-induced mucosal swelling. Fourth, SHS could decrease patency and mucociliary clearance by causing more frequent viral upper respiratory infections. The literature considered in the 2006 report of the Surgeon General included 61 reports based on 59 studies, covering multiple outcomes including acute OM, recurrent OM, middle ear disease, and adenotonsillectomy. The pooled evidence is shown in Fig. 19.6.

FIGURE 19.6 Odds ratios for the effect of smoking by either parent on middle-ear disease in children (U.S. Department of Health and Human Services, 2006).

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19.3 HEALTH EFFECTS OF INVOLUNTARY SMOKING IN ADULTS 19.3.1

Lung Cancer

In 1981, published reports from Japan (Hirayama, 1981) and Greece (Trichopoulos et al., 1981) indicated increased lung cancer risk in nonsmoking women married to cigarette smokers. Subsequently, this association has been examined in over 50 investigations conducted in the United States and other countries. A causal association of involuntary smoking with lung cancer derives biological plausibility from the presence of carcinogens in sidestream smoke and the lack of a documented threshold dose for respiratory carcinogens in active smokers (International Agency for Research on Cancer, 1986; U.S. Department of Health and Human Services, 1982, 1986, 2004). Moreover, genotoxic activity had been demonstrated for many components of SHS (Bennett et al., 1999; Claxton et al., 1989; DeMarini, 2004; Lofroth, 1989; Weiss, 1989). Experimental and real-world exposures of nonsmokers to SHS leads to their excreting NNAL, a tobacco-specific carcinogen, in their urine (Carmella et al., 2003; Hecht et al., 1993). Nonsmokers exposed to SHS also have increased concentrations of adducts of tobacco-related carcinogens (Crawford et al., 1994; Maclure et al., 1989). Additionally, Mauderly et al. (2004), using an animal model, found that whole-body exposure in rats to cigarette smoke increases the risk of neoplastic proliferative lung lesions and induces lung cancer. Time trends of lung cancer mortality in nonsmokers have been examined with the rationale that temporally increasing exposure to SHS should be paralleled by increasing mortality rates (Enstrom, 1979; Garfinkel, 1981). These data provide only indirect evidence on the lung cancer risk associated with involuntary exposure to tobacco smoke. Epidemiologists have directly tested the association between lung cancer and involuntary smoking utilizing conventional designs, the case-control and cohort studies. In a casecontrol study, the exposures of nonsmoking persons with lung cancer to SHS are compared to those of an appropriate control group. In a cohort study, the occurrence of lung cancer over time in nonsmokers is assessed in relation to involuntary tobacco smoke exposure. The results of both study designs may be affected by inaccurate assessment of exposure to SHS, by inaccurate information on personal smoking habits that leads to classification of smokers as nonsmokers, by failure to assess and control for potential confounding factors, and by the misdiagnosis of a cancer at another site as a primary cancer of the lung. As stated previously, methodological investigations suggest that accurate information can be obtained by interview in an epidemiological study on the smoking habits of a spouse (i.e., never or ever smoker) (Coultas et al., 1989; Cummings et al., 1989; Lubin, 1999; Pron et al., 1988). However, information concerning quantitative aspects of the spouse’s smoking is reported with less accuracy. Misclassification of current or former smokers as never smokers may introduce a positive bias because of the concordance of spouse smoking habits (Lee, 1998). The extent to which this bias explains the numerous reports of association between spouse smoking and lung cancer has been addressed, and findings indicate that bias does not account for the observed association (Lee, 1988; U.S. Department of Health and Human Services, 2006; U.S. Environmental Protection Agency, 1992; Wald et al., 1986; Wu, 1999). Use of spouse smoking alone to represent exposure to SHS does not cover exposures outside the home (Friedman et al., 1983) or necessarily all exposure inside the home, particularly during the period relevant to the epidemiological studies. Klepeis et al. (2001)

HEALTH EFFECTS OF INVOLUNTARY SMOKING IN ADULTS

723

used data from the National Human Activity Pattern Survey to assess the contribution of the home and other indoor environments to SHS exposures. Overall, the data show that 43% of the time spent with a smoker is in a residence, while 7% is in the workplace, 9% in a vehicle, and 15% in a bar or restaurant. This survey may help to explain the results of the International Agency for Research on Cancer (IARC), which found that the number of cigarettes smoked per day by the husband is only moderately correlated with “actual” exposure of women married to smokers (Saracci and Riboli, 1989). A subsequent IARC study published in 2004 conducted a pooled analysis of data to assess the risk of lung cancer to nonsmokers exposed to spouse and workplace sources of SHS. They found an excess risk of 23% from exposure to spousal smoking and 27% from exposure to workplace sources of SHS (Brennan et al., 2004). A study in the United States examined the contribution of spouse smoking to total exposure to SHS received at home (Sandler et al., 1989b). Using 1963 data from the Washington County (Maryland) study, Sandler et al. found that for nonsmoking women, spouse smoking contributed 88% of the exposure, whereas for nonsmoking men spouse smoking contributed 62% of the exposure. In some countries, including the United States, smoking prevalence varies markedly with indicators of income and education, more recently tending to rise sharply with decreasing educational level and income (U.S. Department of Health and Human Services, 1989, 2004). In general, exposure to SHS follows a similar trend, and critics of the findings on SHS and lung cancer have argued that uncontrolled confounding by lifestyle, occupation, or other factors may explain the association. In fact, current data for the United States do indicate a generally less healthy lifestyle in those with greater SHS exposure (Matanoski et al., 1995). However, other than a few occupational exposures at high levels, as well as indoor radon, risk factors for lung cancer in never smokers that might confound the SHS association cannot be proffered, and the relevance to past studies of these current associations of potential confounders with SHS exposure is uncertain. The first major studies on SHS and lung cancer were reported in 1981. An early report by Hirayama (1981) was based on a prospective cohort study of 91,540 nonsmoking women in Japan. Standardized mortality ratios (SMRs) for lung cancer increased significantly with the amount smoked by the husbands. The findings could not be explained by confounding factors and were unchanged when follow-up of the study group was extended (Hirayama, 1984). On the basis of the same cohort, Hirayama (1984) also reported significantly increased risk for nonsmoking men married to wives smoking 1–19 cigarettes and 20 or more cigarettes daily. In 1981, Trichopoulos et al. (1981) also reported increased lung cancer risk in nonsmoking women married to cigarette smokers. These investigators conducted a case-control study in Athens, Greece, which included cases with a diagnosis other than for orthopedic disorders. The positive findings reported in 1981 were unchanged with subsequent expansion of the study population (Trichopoulos et al., 1983). By 1986, the evidence had mounted, and three reports published in that year concluded that SHS was a cause of lung cancer. The IARC of the World Health Organization (International Agency for Research on Cancer, 1986) concluded that “passive smoking gives rise to some risk of cancer.” In its monograph on tobacco smoking, the agency supported this conclusion on the basis of the characteristics of sidestream and mainstream smoke, the absorption of tobacco smoke materials during involuntary smoking, and the nature of dose–response relationships for carcinogenesis. In the same year, the NRC (National Research Council and Committee on Passive Smoking, 1986) and the U.S.

724

SECONDHAND SMOKE

Surgeon General (U.S. Department of Health and Human Services, 1986) also concluded that involuntary smoking increases the incidence of lung cancer in nonsmokers. In reaching this conclusion, the NRC cited the biological plausibility of the association between exposure to SHS and lung cancer and the supporting epidemiological evidence. Based on a pooled analysis of the epidemiological data adjusted for bias, the report concluded that the best estimate for the excess risk of lung cancer in nonsmokers married to smokers was 25%. The 1986 report of the Surgeon General (U.S. Department of Health and Human Services, 1986) characterized involuntary smoking as a cause of lung cancer in nonsmokers. This conclusion was based on the extensive information already available on the carcinogenicity of active smoking, on the qualitative similarities between SHS and mainstream smoke, and on the epidemiological data on involuntary smoking. In 1992, the EPA (U.S. Environmental Protection Agency, 1992) published its risk assessment of SHS as a Group A carcinogen. The agency’s evaluation drew on the toxicologic evidence on SHS and the extensive literature on active smoking. A metaanalysis of the 31 studies published to that time was central in the decision to classify SHS as a Group A carcinogen—namely, a known human carcinogen. The meta-analysis considered the data from the epidemiologic studies by tiers of study quality and location and used an adjustment method for misclassification of smokers as never smokers. Overall, the analysis found a significantly increased risk of lung cancer in never-smoking women married to smoking men; for the studies conducted in the United States, the estimated relative risk was 1.19 (90% CI: 1.04, 1.35). Critics of the report have raised a number of concerns including the use of meta-analysis, reliance of 90% rather than 95% confidence intervals, uncontrolled confounding, and information bias. The report, however, was endorsed by the Agency’s Science Advisory Board, and its conclusion is fully consistent with the 1986 reports. Subsequent to the 1992 risk assessment, over a dozen additional studies and three major reports have been published that further contribute to the evidence supporting a causal association between secondhand smoke and the risk of lung cancer (California Environmental Protection Agency and Office of Environmental Health Hazard Assessment, 1997; Cardenas et al., 1997; Fontham et al., 1994; International Agency for Research on Cancer, 2004; Kabat et al., 1995; U.S. Department of Health and Human Services, 2006). Among the additional studies, the multicenter study of Fontham et al. (1994) is one of the largest to date, with 651 cases and 1253 controls. It shows a significant increase in overall relative risk (OR ¼ 1.26, 95% CI ¼ 1.04, 1.54). Significant risk was also associated with occupational exposure to SHS. Findings of an autopsy study conducted in Greece also strengthened the plausibility of the lung cancer/SHS association. Trichopoulos et al. (1992) examined autopsy lung specimens from 400 persons 35 years of age and older to assess airways changes. Epithelial lesions were more common in nonsmokers married to smokers than in nonsmokers married to nonsmokers. Hackshaw et al. (1997) carried out a comprehensive meta-analysis that included 37 published studies. They estimated an excess risk of lung cancer for smokers married to nonsmokers as 24% (95% CI: 13%, 36%). Adjustment for potential bias and confounding by diet did not alter the estimate. This meta-analysis was part of the basis for the conclusion by the U.K. Scientific Committee on Tobacco and Health (Scientific Committee on Tobacco and Health and HSMO, 1998) that SHS is a cause of lung cancer. A subsequent IARC (2004) meta-analysis including 46 studies and 6257 cases yielded similar results, 24% (95% CI: 14%, 34%), and incorporating the results from a cohort study with null results overall,


725

but only 177 cases (Enstrom and Kabat, 2003), did not change the findings (Hackshaw, 2003). The most recent summaries from the 2006 Surgeon General’s report are provided in Table 19.3. The extent of the lung cancer hazard associated with involuntary smoking in the United States and in other countries remains subject to some uncertainty, however, although estimates have been made that are useful indications of the magnitude of the disease risk (U.S. Department of Health and Human Services, 1986; Weiss, 1986). Risk estimation procedures have been used to describe the lung cancer risk associated with involuntary smoking, but assumptions and simplifications are necessary to apply this method. The estimates of lung cancer deaths attributable to passive smoking have previously received widespread media attention and have figured prominently in the evolution of public policy on passive smoking. In 1990, Repace and Lowrey (1990) reviewed the risk assessments of lung cancer and passive smoking and estimated the numbers of lung cancer cases in U.S. nonsmokers attributable to passive smoking. The range of the nine estimates, covering both never smokers and former smokers, provided by Repace and Lowery was from 58 to 8124 lung cancer deaths for the year 1988, with an overall mean of 4500 or 5000 excluding the lowest estimate of 58. The bases for the individual estimates included the comparative dosimetry of tobacco smoke in smokers and nonsmokers using presumed inhaled dose or levels of nicotine or cotinine, the epidemiological evidence, and modeling approaches. The 1992 estimate of the EPA, based on the epidemiologic data was about 3000, including 1500 and 500 deaths in never-smoking women and men, respectively, and about 100 in long-term former smokers of both sexes (U.S. Environmental Protection Agency, 1992). More recently, Repace et al. (1998) developed a model of risk to workers of lung cancer and heart disease arising from SHS exposure. The pharmacokinetic model incorporated nicotine as an indicator of exposure and cotinine as a measure of dose to estimate risks. The model estimated that 400 lung cancer deaths occur annually from workplace exposure at a prevalence of 28% smoking in the workplace. The California EPA estimates that at least 3423, and perhaps as many as 8866, lung cancer deaths were caused by SHS across the nation in 2003 alone. Of those 3423 deaths, 967 were due to nonspousal exposures to secondhand smoke and 2456 were due to spousal exposure (California Environmental Protection Agency and Air Resources Board, 2005). These calculations illustrate that passive smoking must be considered an important cause of lung cancer death from a public health perspective; exposure is involuntary and not subject to control. The specific risk assessments require assumptions concerning the extent and degree of exposure to SHS, exposure–response relationships, and the lifetime expression of the excess risk associated with passive smoking at different ages. Moreover, the calculations did not consider the potential contributions of other exposures, such as occupational agents and indoor radon. The current decline in the prevalence of active smoking, and the implementation of strong clean indoor air policies, will reduce the relevance of estimates based on past patterns of smoking behavior. 19.3.2

Other Cancers

In adults, active smoking has been linked to a generally increased risk of malignancy and to excess risk at specific sites, including lung, urinary tract, upper aerodigestive tract, liver, stomach, pancreas, and others (U.S. Department of Health and Human Services, 2004; Vineis et al., 2004), reflecting the widespread effects of carcinogens in tobacco smoke. Given

726

SECONDHAND SMOKE

TABLE 19.3 Quantitative Estimate of Lung Cancer with Differing Sources of Exposure to Secondhand Smoke (U.S. Department of Health and Human Services, 2006) Study Hackshaw (1997) Zhong (2000)

Data Source 37 studies

Surgeon General report 2006 Spouse

40 studies (including 37 from Hackshaw study) Case control (448 studies) Cohort (8 studies)

54 studies

Men

Women

United States and Canada Europe Asia Surgeon General report 2006 Workplace 25 studies

Surgeon General report 2006 Childhood 24 studies

Nonsmokers (25 studies) Nonsmoking men (11 studies) Nonsmoking women (25 studies) Nonsmokers United States and Canada (8 studies) Nonsmokers Europe (7 studies) Nonsmokers Asia (10 studies) Men and women Men and women Men and women Women Women United States (8 studies) Europe (6 studies) Asia (10 studies)

Exposure Versus Referent Smoking versus nonsmoking spouse Smoking husband versus nonsmoking husband Smoking versus nonsmoking spouse Smoking versus nonsmoking spouse Smoking wife versus nonsmoking wife Smoking husband versus nonsmoking husband Smoking versus nonsmoking spouse Smoking versus nonsmoking spouse Smoking versus nonsmoking spouse Workplace SHS versus not Workplace SHS versus not Workplace SHS versus not Workplace SHS versus not Workplace SHS versus not Workplace SHS versus not Maternal smoking Paternal smoking Either parent smoking Maternal smoking Paternal smoking Either parent smoking Either parent smoking Either parent smoking

RR

95% CI

1.24

1.13–1.36

1.20

1.12–1.29

1.21

1.13–1.30

1.29

1.125–1.49

1.37

1.05–1.79

1.22

1.13–1.31

1.15

1.04–1.26

1.16

1.03–1.30

1.43

1.24–1.66

1.22

1.13–1.33

1.12

0.86–1.50

1.22

1.10–1.35

1.24

1.03–1.49

1.13

0.96–1.34

1.32

1.13–1.55

1.15

0.86–1.52

1.10 1.11

0.89–1.36 0.94–1.31

1.28 1.17 0.93

0.93–1.78 0.91–1.50 0.81–1.07

0.81

0.71–0.92

1.59

1.18–2.15


727

that the same carcinogens are present in both mainstream and sidestream smoke and that these carcinogens have no evident threshold level in active smokers and have a demonstrated uptake by involuntary smokers, there is a compelling rationale for the hypothesis that secondhand smoke exposure increases the risk of cancers that are also caused by active smoking. While there is substantial and sufficient epidemiologic evidence for a causal association between SHS exposure and lung cancer, less data are available to analyze the causal association for other cancer sites. However, the data are mounting for some sites, and controversy has ensued, particularly for the association between SHS exposure and breast cancer. Two early studies assessed the association between SHS exposure and cancer in general (Miller, 1984; Sandler et al., 1985a, 1985b, 1985c). Miller (1984) interviewed surviving relatives of 537 deceased nonsmoking women in western Pennsylvania concerning the smoking habits of their husbands. A significantly increased risk of cancer death (OR ¼ 1.94; P < 0.05) was found in women who were married to smokers and also not employed outside the home. The large number of potential subjects who were not interviewed and the possibility of information bias detract from this report. Sandler et al. (1985a, 1985b, 1985c) conducted a case-control study on the effects of exposure to SHS during childhood and adulthood on the risk of cancer. The 518 cases included cancers of all types other than basal cell cancer of the skin; the cases and the matched controls were between the ages of 15 and 59 years. For all sites combined, significantly increased risk was found for parental smoking (crude OR ¼ 1.6), and for marriage to a smoking spouse (crude OR ¼ 1.5); the effects of these two exposures were independent (Sandler et al., 1985c). Significant associations were also found for some individual sites: for childhood exposure (Sandler et al., 1985b), maternal and paternal smoking increased the risk of hematopoietic malignancy, and for adulthood exposure (Sandler et al., 1985a), spouse’s smoking increased the risk for cancers of the female breast, female genital system, and the endocrine system. The findings are primarily hypothesis generating and require replication. In a case-control study, such as those reported by Sandler et al., information on exposure to SHS may be affected by information bias. One cancer site of particular interest at present is breast cancer. Given the widespread exposure to secondhand smoke, this exposure could potentially be an important avoidable cause of breast cancer. In considering whether passive smoking causes breast cancer, the evidence for active smoking needs to be considered in assessing the plausibility of an association of breast cancer risk with secondhand smoke in nonsmokers. There is some evidence to suggest that an association between tobacco smoke and breast cancer is biologically plausible. Studies have shown that carcinogens in tobacco smoke reach breast tissue (Li et al., 1996; Petrakis et al., 1978, 1988) and are mammary mutagens (Dunnick et al., 1995; El Bayoumy et al., 1995; Nagao et al., 1994). However, other studies using biomarkers have found an association between smoking and decreased levels of estrogen (MacMahon et al., 1982; Michnovicz et al., 1986), which implies that active smoking might decrease risk of breast cancer. Furthermore, the 2001 and 2004 reports of the Surgeon General found that smoking was associated with a decreased risk of endometrial cancer and an earlier age at menopause (U.S. Department of Health and Human Services, 2001, 2004). These antiestrogenic consequences of active smoking have been construed as implying that breast cancer risk would be reduced for active smokers in comparison to never smokers. The evidence is not consistent, however, and uncertainty remains about the effect of smoking on blood estrogen levels. These possibly opposing biological consequences of active smoking may explain why

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review of the epidemiologic data has found an overall null effect of active smoking on the risk of breast cancer. Since the 1960s, there have been more than 50 studies investigating the association between active smoking and breast cancer. In 2002, Hamajima et al. (2002) conducted a pooled analysis of data from 53 studies and found a relative risk of 0.99 (95% CI: 0.92, 1.05) for women who were current smokers compared with women who were lifetime nonsmokers. One possible explanation for the null results in Hamajima’s pooled analysis is that the antiestrogenic effects of smoking may offset the potentially carcinogenic effects on the risk of breast cancer. Subsequently, the 2004 reports of the Surgeon General and of IARC concluded that the weight of evidence strongly suggests that there is no causal association between active smoking and breast cancer (U.S. Department of Health and Human Services, 2004; International Agency for Research on Cancer, 2004). One year later, the California EPA concluded that active smoking is a cause of breast cancer, although it did not carry out a full systematic review (California Environmental Protection Agency and Air Resources Board, 2005). Two cohort studies published in 2004 found a significant increase in risk of breast cancer (Al Delaimy et al., 2004; Reynolds et al., 2004). However, sufficient evidence has not accumulated to suggest a causal association between active smoking and breast cancer (U.S. Department of Health and Human Services, 2006). More than 20 epidemiologic studies have been published specifically addressing the association between secondhand smoke and breast cancer. Several major reports, including the IARC report, the California EPA 2005 report, and the Surgeon General 2006 report, have reviewed the evidence for an association between SHS exposure and breast cancer (California Environmental Protection Agency and Air Resources Board, 2005; International Agency for Research on Cancer, 2004; U.S. Department of Health and Human Services, 2006). The Cal/EPA conducted a meta-analysis using six cohort studies and 12 case-control studies that were deemed to provide the “best evidence.” They found an increased risk of 25% (95% CI: 8%, 44%) overall and concluded that there is sufficient evidence for a causal association among premenopausal women (California Environmental Protection Agency and Air Resources Board, 2005). Among postmenopausal women, there was no indication of an association. In 2004, the IARC concluded that the evidence is inconsistent, and although some case-control studies found positive effects, cohort studies overall did not find a causal association (International Agency for Research on Cancer, 2004). Additionally, the lack of a positive dose–response relationship and the lack of association with active smoking weigh against the possibility of an increased risk of breast cancer from SHS exposure. Subsequently, the Surgeon General came to similar conclusions (U.S. Department of Health and Human Services, 2006). Using data from seven prospective cohort studies and 14 casecontrol studies, the Surgeon General’s report conducted a meta-analysis. Sensitivity analyses showed that cohort studies overall found null results and studies that adjusted for potential confounding showed weaker associations (U.S. Department of Health and Human Services, 2006). Furthermore, the Surgeon General’s report evaluated the possibility of publication bias and found that less precise studies tended to have more positive results. Finally, after reviewing all the evidence using the criteria for causality, the Surgeon General’s report found that overall the evidence is inconsistent and concluded that the evidence is suggestive but not sufficient to infer a causal association between SHS exposure and breast cancer. Other studies provide data on passive smoking and cancers of diverse sites. Hirayama (1984) has reported significantly increased mortality from nasal sinus cancers and from brain tumors in nonsmoking women married to smokers in the Japanese cohort. Additionally, two


729

case-control studies, one in Japan and one in the United States, found an association between nasal sinus cancer and SHS exposure in nonsmokers living with a smoking spouse (Fukuda and Shibata, 1990; Zheng et al., 1993). However, these case-control studies relied on relatively small sample sizes, using fewer than 170 cases. Nasopharyngeal carcinoma has also been investigated as a potential cancer caused by SHS exposure in three published case-control studies (Cheng et al., 1999; Yu et al., 1990; Yuan et al., 2000). Two of the three studies found null effects among nonsmokers (Cheng et al., 1999; Yu et al., 1990), and the third found an association among women, but nonsignificant effects in men (Yuan et al., 2000). The 2006 Surgeon General’s report concluded that “the evidence is inadequate to infer the presence or absence of a causal relationship” for nasopharyngeal carcinoma (U.S. Department of Health and Human Services, 2006). Cervical cancer, which has been linked to active smoking (U.S. Department of Health and Human Services, 1990b), was associated with duration of involuntary smoking in a casecontrol study in Utah (Slattery et al., 1989) and with cumulative exposure in a case-control study by Sandler et al. (1985c), looking at only 62 cases of cervical cancer. In 2003, Wu et al. found an increased risk of cervical intraepithelial neoplasia among nonsmoking women exposed to SHS exposure at home, suggesting a possible role for secondhand smoke in the etiology of cervical cancer (Wu et al., 2003). Two other case-control studies found borderline statistical significance for an association between SHS exposure and cervical cancer (Coker et al., 1992) and for an association between SHS exposure and abnormal Pap smear results (Scholes et al., 1999). Among the two published cohort studies investigating an association between cervical cancer and SHS exposure in nonsmoking wives married to smokers, Hirayama et al. (Hirayama, 1981) found a nonsignificant increased risk in a Japanese cohort and Jee et al. (2002) found no association in a Korean cohort. Based on a review of the evidence, the 2006 Surgeon General’s report concluded that a causal association between SHS exposure and cervical cancer cannot yet be inferred (U.S. Department of Health and Human Services, 2006). In a case-control study of bladder cancer, involuntary smoke exposure at home and at work did not increase risk (Kabat et al., 1986). Another case-control study by Burch et al. also found null effects for the association between SHS exposure and bladder cancer (Burch et al., 1989). In 2002, Zeegers et al. found a slight but nonsignificant increase in risk of bladder cancer in nonsmokers exposed in the workplace (Zeegers et al., 2002). Although the association between bladder cancer and active smoking has been established (International Agency for Research on Cancer, 1986), there is insufficient evidence to conclude that SHS exposure is causally linked to bladder cancer. In the Washington County (Maryland) study, colorectal cancer incidence rates were significantly increased for male secondhand smokers, but not for female secondhand smokers; incidence rates were significantly reduced for female active smokers (Sandler et al., 1988). This pattern of findings cannot be readily explained. These associations of involuntary smoking with cancer at diverse nonrespiratory sites cannot be readily supported with arguments for biological plausibility. Increased risks at some of the sites, such as female breast cancer, have not been observed in active smokers (International Agency for Research on Cancer, 2004; U.S. Department of Health and Human Services, 1989, 1990a, 2004, 2006). In fact, the IARC has concluded that effects would not be produced in passive smokers that would not be produced to a larger extent in active smokers (International Agency for Research on Cancer, 1986, 2004). Thus, investigation of cancer sites other than the lung should be guided by the data from active smokers and by appropriate toxicological evidence. For example, the plausibility of secondhand smoking

730

SECONDHAND SMOKE

with cervical cancer would be supported by the demonstration of tobacco smoke components in the cervical mucus of exposed nonsmoking women. In investigations of cancer at sites not plausibly linked to secondhand smoke exposure, associations may arise by chance or by the effect of bias. Amassing data on secondhand smoke and all cancers may thus produce noncausal associations, prompting further but possibly unnecessary investigations.

19.4 SHS AND CORONARY HEART DISEASE 19.4.1

Introduction

Causal associations between active smoking and fatal and nonfatal coronary heart disease (CHD) outcomes have long been demonstrated (U.S. Department of Health and Human Services, 2004). This increased risk of CHD morbidity and mortality has been found in younger persons and the elderly, men and women, and ethnically and racially diverse populations. The risk of CHD in active smokers increases with amount and duration of cigarette smoking and decreases during the first year after cessation. Active cigarette smoking is considered to increase the risk of cardiovascular disease by promoting atherosclerosis, affecting endothelial cell functioning, increasing the tendency to thrombosis, causing spasm of the coronary arteries that increases the likelihood of cardiac arrhythmias, and decreasing the oxygen-carrying capacity of the blood (U.S. Department of Health and Human Services, 1990b). It is biologically plausible that exposure to secondhand smoke could also be associated with increased risk for CHD through the same mechanisms considered relevant for active smoking, although the lower exposures to smoke components of the secondhand smoker have raised questions regarding the relevance of the mechanisms cited for active smoking. 19.4.2

Biological Plausibility

In 2005, Barnoya and Glantz (2005) summarized the pathophysiological mechanisms by which secondhand smoke exposure might increase the risk of heart disease. They suggest that passive smoking may promote atherogenesis, increase the tendency of platelets to aggregate and thereby promote thrombosis, impair endothelial cell function, increase arterial stiffness leading to atherosclerosis, reduce the oxygen-carrying capacity of the blood, and alter myocardial metabolism, much as for active smoking and CHD. Several separate experiments involving exposure of nonsmokers to SHS have shown that passive smoking affects measures of platelet function in the direction of increased tendency toward thrombosis (Barnoya and Glantz, 2005; Glantz and Parmley, 1995). In a 2004 study by Rubenstein et al., sidestream smoke was found to be 50% more potent than mainstream smoke in activating platelets (Rubenstein et al., 2004). Glantz and Parmley also proposed that carcinogenic agents such as polycyclic aromatic hydrocarbons found in tobacco smoke promote atherogenesis by effects on cell proliferation (Glantz and Parmley, 1995). Exposure to secondhand smoke may also worsen the outcome of an ischemic event in the heart: animal data have demonstrated that SHS exposure increases cardiac damage following an experimental myocardial infarction. Experiments on two species of animals (rabbits and cockerels) have demonstrated that not only does exposure to SHS at doses similar to exposure to humans accelerate the growth of atherosclerotic plaques through the increase of lipid deposits, but it also induces atherosclerosis. There is also impressive and accumulating evidence that SHS

SHS AND CORONARY HEART DISEASE

731

affects vascular endothelial cell functioning (Celermajer et al., 1996; Otsuka et al., 2001; Sumida et al., 1998). Otsuka et al. found that 30 min of exposure to secondhand smoke in healthy young volunteers compromised coronary artery endothelial function in a manner that was indistinguishable from that of habitual smokers, suggesting that endothelial dysfunction may be an important mechanism by which exposure to secondhand smoke increases CHD risk (Otsuka et al., 2001). In addition to its effects on platelets, SHS exposure affects the oxygen-carrying capacity of the blood. Even small increments, on the order of 1%, in the carboxyhemoglobin, may explain finding that secondhand smoke exposure decreases the duration of exercise of patients with angina pectoris (Allred et al., 1989). This is supported with evidence that cigarette smoking has been shown to increase levels of CO in the spaces where ventilation is low or smoking is particularly intense (U.S. Department of Health and Human Services, 1986). 19.4.3

Epidemiological Studies

Epidemiologic data first raised concern that exposure to secondhand smoke may increase risk for CHD with the 1985 report of Garland et al. (1985) based on a cohort study in southern California. There are now more than 20 studies on the association between secondhand smoke and cardiovascular disease, including 11 cohort and 12 case-control studies, and 1 cross-sectional study. These studies assessed both fatal and nonfatal cardiovascular heart disease outcomes, and most used self-administered questionnaires to assess SHS exposure. They cover a wide range of populations, both geographically and racially. While many of the studies were conducted within the United States, studies were also conducted in Europe (Scotland, Italy, United Kingdom, Sweden), Asia (Japan and China), South America (Argentina), and the South Pacific (Australia and New Zealand). The majority of the studies measured the effect of SHS exposure due to spousal smoking; however, some studies also assessed exposures from smoking by other household members or occurring at work or in transit. Several studies included measurement of biomarkers. One group of studies addresses the promotion of atherosclerosis and SHS exposure, particularly through the mechanism of increased carotid intimal medial thickness (IMT). Three studies that assessed the link between SHS and carotid IMT found an increase in carotid IMT with exposure to SHS of nonsmokers (Diez-Roux et al., 1995; Howard et al., 1994). Using data from the ARIC Study, Howard et al. found an average difference in carotid IMT of 13 mm between exposed and unexposed nonsmokers after adjusting for possible confounders (Howard et al., 1994). They also found a dose–response relationship between exposure to SHS and increased carotid IMT in male nonsmokers. Diez-Roux and colleagues also found an increase in carotid IMTwith exposure to SHS, analyzing data from 2073 ARIC participants who were lifetime nonsmokers. The participants were assessed on two separate occasions in 1975 and 1987–1989. Regardless of whether they were exposed to SHS on the first, second, or both occasions, there was a significant increase in carotid IMT at final assessment (Diez-Roux et al., 1995). These findings suggest that SHS has a long-term effect on carotid IMT and provide further evidence for the association between SHS and atherosclerosis. Subsequently, in 1998 Howard et al. examined the effect of SHS exposure on progression of IMT. They found that nonsmokers exposed to SHS had a 20% increased IMT progression rate compared to those unexposed (Howard et al., 1998). As the evidence has subsequently mounted since the Garland et al. report, it has been reviewed systematically by the American Heart Association (Taylor et al., 1992), the

732

SECONDHAND SMOKE

California EPA (California Environmental Protection Agency and Air Resources Board, 2005; California Environmental Protection Agency and Office of Environmental Health Hazard Assessment, 1997), the Scientific Committee on Tobacco and Health in the United Kingdom (Scientific Committee on Tobacco and Health and HSMO, 1998) and most recently by the U.S. Surgeon General (U.S. Department of Health and Human Services, 2006). Review of the evidence has uniformly led to the conclusion that there is a causal association between exposure to secondhand smoke and risk of cardiovascular disease (California Environmental Protection Agency and Office of Environmental Health Hazard Assessment, 1997; Scientific Committee on Tobacco and Health and HSMO, 1998). The meta-analysis prepared for the 2006 U.S. Surgeon General’s Report, including nine cohort studies and seven case-control studies, estimated the pooled excess risk from SHS exposures as 27% (95% CI: 19–36%) (U.S. Department of Health and Human Services, 2006). The forest plot in Fig. 19.7 summarizes the evidence reviewed in the 2006 Surgeon

FIGURE 19.7 Pooled relative risks of coronary heart disease (CHD) associated with secondhand smoke exposure among nonsmokers in various subgroups (U.S. Department of Health and Human Services, 2006).

SHS AND CORONARY HEART DISEASE

733

General’s report (U.S. Department of Health and Human Services, 2006). Among the cohort studies included in the meta-analysis, seven were conducted in the United States, one in Japan, and one in Scotland. Among the case-control studies, one was conducted in the United States and the others were conducted abroad. Nine of the studies, both cohort and casecontrol, included men and women subjects, while six included only women. Only one study was restricted to men. In all studies, participants were nonsmokers and except for three of the studies, they were lifetime nonsmokers. For those three studies, either the study did not explicitly indicate that participants had formerly smoked or it included former smokers (Butler, 1988; Hirayama, 1984; McElduff et al., 1998). In all but one study, investigators assessed exposure to secondhand smoke in the home from either a spouse or cohabitant through self-report. Four of the studies also assessed exposure to secondhand smoke outside of the home, including workplace exposures. Two of the studies did not specify or distinguish among the different sources of secondhand smoke exposure. Among the cohort studies reporting on the effect of exposure to secondhand smoke on cardiovascular outcomes, five of the cohort studies used fatal CHD as the primary outcome, three used ischemic heart disease (IHD), and one combined fatal CHD and nonfatal acute myocardial infarction. Among the case-control studies, four used nonfatal acute myocardial infarction as the primary outcome, one used nonfatal CHD, one used fatal and nonfatal acute myocardial infarction, and one used nonfatal IHD. While the risk estimates for SHS and CHD outcomes vary in these studies, they range mostly from null to modestly significant increases in risk, with the risk for fatal outcomes generally higher and more significant. Additionally, a prospective cohort study reported in 2004 used serum cotinine levels for exposure classification (Whincup et al., 2004). The study included 4729 men in the United Kingdom who provided baseline blood samples in 1978–1980. After 20 years of follow-up, among the 2105 men who were nonsmokers, the risk of CHD was increased in those with higher serum cotinine concentrations. Compared to men in the lowest quartile of serum cotinine concentration, after adjusting for established CHD risk factors, the risks in the second, third, and fourth quartiles were 1.45, 1.49, and 1.57, respectively. A consistent association was not found between serum cotinine concentration and stroke. However, there is increasing epidemiologic evidence suggestive of a causal association between SHS exposure and stroke. At least seven epidemiologic studies (four case-control, one cohort, and two cross-sectional studies) have been published exploring the association between SHS exposure and stroke (Bonita et al., 1999; Donnan et al., 1989; Howard et al., 1998; Lee et al., 1986; Sandler et al., 1989a; You et al., 1999; Zhang et al., 2005). A large cross-sectional study of 60,377 women in China found an association between prevalent stroke in women and smoking by their husbands (Zhang et al., 2005). The prevalence of stroke increased with greater duration of smoking and with an increasing number of cigarettes smoked daily. Sandler et al. conducted a cohort study of 19,035 lifetime nonsmokers using census data from Washington County, MD. Based on 297 cases among women exposed to SHS, they found a 24% increased risk of stroke compared with unexposed. They found null results for an association in men but were limited to only 33 cases among men (Sandler et al., 1989a). A case-control study in New Zealand did find a twofold increased risk of stroke in men exposed to SHS, looking at 215 cases and 1336 controls (Bonita et al., 1999). 19.4.4

Conclusions

There are strengths and weaknesses to both the case-control and cohort study designs in investigating SHS and CHD outcomes. Many of the case-control studies suffer from

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small sample sizes and lack the power to detect significant associations. Furthermore, many studies also lack information on other risk factors for CHD, and therefore they do not adequately adjust for confounders. In contrast, many of the cohort studies have large sample sizes and do adjust for confounders. They also avoid information bias by assessing smoking status and exposure prior to the CHD outcome. However, cohort studies are more susceptible to exposure misclassification due to the cessation or resumption of smoking by the source of exposure; this risk of misclassification increases with the length of follow-up. Although the risk estimates for SHS and CHD outcomes vary, they range mostly from null to modestly significant increases in risk, with the risk for fatal outcomes generally higher and more significant. In their meta-analysis, Law et al. (1997) estimated the excess risk from SHS exposure as 30% (95% CI: 22–38%) at age 65 years. In 1997, the California Environmental Protection Agency (California Environmental Protection Agency and Office of Environmental Health Hazard Assessment, 1997) concluded that there is “an overall risk of 30%” for CHD due to exposure from SHS (California Environmental Protection Agency and Office of Environmental Health Hazard Assessment, 1997). In 2005, Cal/EPA established that 22,700–69,500 deaths from CHD were attributable to SHS in 2000 (California Environmental Protection Agency and Air Resources Board, 2005). The American Heart Association’s Council on Cardiopulmonary and Critical Care has also concluded that secondhand smoke both increases the risk of heart disease and is “a major preventable cause of cardiovascular disease and death” (Taylor et al., 1992). This conclusion was echoed in 1998 by the Scientific Committee on Tobacco and Health, both Cal/EPA reports, and the 2006 report of the Surgeon General (California Environmental Protection Agency and Office of Environmental Health Hazard Assessment, 1997; Scientific Committee on Tobacco and Health and HSMO, 1998; California Environmental Protection Agency and Air Resources Board, 2005; U.S. Department of Health and Human Services, 2006).

19.5 RESPIRATORY SYMPTOMS AND ILLNESSES IN ADULTS Extensive evidence has shown that active smoking causes respiratory symptoms and illnesses (U.S. Department of Health and Human Services, 2004). Active smoking can cause inflammatory injury throughout the respiratory tract, leading to both acute and chronic respiratory symptoms, impaired lung function, and eventually to chronic obstructive pulmonary disease (COPD). The similarity of mainstream and SHS implies that involuntary smoking might also cause inflammation of the respiratory tract. One autopsy study of the lungs of nonsmokers found inflammation associated with SHS exposure (Trichopoulos et al., 1992). The 2006 report of the Surgeon General concluded that there are multiple mechanisms by which SHS exposure may cause respiratory symptoms and illnesses (U.S. Department of Health and Human Services, 2006). Observational studies provide evidence that SHS exposure can cause respiratory effects in adults. In 1986, both the NRC and the Surgeon General’s reports not only concluded that SHS exposure is an irritant (National Research Council and Committee on Passive Smoking, 1986; U.S. Department of Health and Human Services, 1986), but also stated that additional evidence was needed to determine if there is a causal association of exposure with chronic respiratory symptoms and reduced pulmonary function. Since then, many studies and several major reports (California Environmental Protection Agency and Air Resources Board, 2005; California Environmental Protection Agency and Office of Environmental Health Hazard Assessment, 1997; U.S. Department of Health and Human Services, 2006; U.S. Environ-

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mental Protection Agency, 1992) have been published that considered the association between SHS exposure and respiratory health in nonsmokers. 19.5.1

Respiratory Symptoms

The association between respiratory symptoms in nonsmokers and involuntary exposure to tobacco smoke has largely been investigated in experimental and cross-sectional studies. The experimental evidence primarily comes from studies that assessed acute responses of asthmatics who were exposed to SHS in a chamber. This experimental approach cannot be readily controlled because of the impossibility of blinding subjects to their being exposed to SHS. However, suggestibility does not appear to underlie physiological responses of asthmatics to SHS (Urch et al., 1988). Of the three studies involving exposure of unselected asthmatics to SHS, only one showed a definite adverse effect (Dahms et al., 1981; Hargreave et al., 1981; Murray and Morrison, 1986; Qin et al., 1991; Shephard et al., 1979). Stankus et al. (1988) recruited 21 asthmatics who reported exacerbation with exposure to SHS. With challenge in an exposure chamber at concentrations much greater than that typically encountered in indoor environments, seven subjects experienced a more than 20% decline in FEV1. Among the 13 epidemiologic studies that were reviewed in the 2006 Surgeon General’s report, only a few were longitudinal in their design (Jaakkola et al., 1996; Robbins et al., 1993; Schwartz and Zeger, 1990). Consistent evidence of an effect of SHS exposure on acute respiratory symptoms in adults has been found (U.S. Department of Health and Human Services, 2006). However, the evidence of an effect of SHS on chronic respiratory symptoms has been less consistent (U.S. Department of Health and Human Services, 2006). Overall, symptoms of chronic cough and dyspnea have been more consistently associated with exposure to SHS than have the symptoms of chronic phlegm and wheeze (U.S. Department of Health and Human Services, 2006). Several studies suggest that exposure to SHS may cause acute respiratory morbidity. Analysis of National Health Interview Survey data showed that a pack-a-day smoker increases respiratory restricted days by about 20% for a nonsmoking spouse (Ostro, 1989). In a study of determinants of daily respiratory symptoms in Los Angeles student nurses, it was found that there was a significantly increased risk of an episode of phlegm with a smoking roommate, after controlling for personal smoking (Schwartz and Zeger, 1990). Several studies have addressed chronic respiratory symptoms. Leuenberger et al. (1994) describe associations between passive exposures to tobacco smoke, at home and in the workplace, and respiratory symptoms in 4197 randomly selected never-smoking adults in the Swiss Study on Air Pollution and Lung Diseases in Adults, a multicenter study in eight areas of the country. Exposed subjects were those who reported any exposure during the past 12 months; exposed persons were then asked about workplace exposure and also about the number of smokers and the duration of exposure at home and work together. Involuntary smoke exposure was associated with asthma, dyspnea, bronchitis and chronic bronchitis symptoms, and allergic rhinitis. The increments in risk were substantial, ranging approximately 40–80% for the different respiratory outcome measures. The increments were not reduced by control for educational level, and dose–response relationships were found with the quantitative indicators of exposure. For several of the outcome measures, the dose–response relationships tended to be steeper for those who also reported workplace exposure. In a cross-sectional population study of 1954 women, Baker and

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Henderson (1999) found a significant association (OR ¼ 1.73; 95% CI: 1.05, 2.85) of wheeze in nonsmoking mothers living with a smoking partner. No association was found in fathers. Other adverse respiratory effects of involuntary smoking have been found in adults. Robbins et al. (1993) examined predictors of new symptoms compatible with “airway obstructive disease” in a cohort study of 3914 nonsmoking participants in the Adventist Health Study. Significantly increased risk was identified in association with exposure during both childhood and adulthood. In a cross-sectional study, Dayal et al. (1994) found that never-smoking Philadelphia residents with a reported diagnosis of asthma, chronic bronchitis, or emphysema had sustained significantly greater exposure to tobacco smoke than unaffected controls. Several epidemiologic studies have investigated the roles of SHS exposure in the onset of asthma and in exacerbating asthma in adults. Two cross-sectional studies, one case-control study, and two prospective cohort studies all found an association between SHS exposure and asthma morbidity (Jindal et al., 1994, 1999; Mannino et al., 1997; Ostro et al., 1994; Sippel et al., 1999; Tarlo et al., 2000). A large cross-sectional study by Mannino et al., using data from the 1991 National Health Interview Survey (Mannino et al., 1997), found that lifetime nonsmokers exposed to SHS had a 44% increased risk of exacerbated chronic respiratory conditions, compared to unexposed nonsmokers, after adjusting for potential confounders. In a small case-control study, Tarlo et al. (2000) found that asthma patients with an exacerbation were more likely to have been exposed to SHS than asthma patients without an exacerbation. In a cohort study of lifetime nonsmoking asthma patients by Jindal et al., SHS exposure was found to increase the risk of acute episodes and impaired lung function (Jindal et al., 1994). Ostro et al. (1994) also found an increased risk of shortness of breath, cough, and restricted activity among asthmatics exposed to SHS compared to unexposed asthmatics. Sippel et al. (1999) found a significant increase in the use of hospital services, such as urgent care and emergency room visits, among asthma patients exposed to SHS compared to those who were unexposed. In a prospective cohort study of adult nonsmokers admitted to the hospital for asthma, Eisner et al. found a significant association between the severity of asthma symptoms and SHS exposure after controlling for potential confounders (Eisner et al., 2005). In 1999, Weiss, Utell and Samet reviewed the literature on SHS exposure and asthma in adults (Weiss et al., 1999). They found that two prospective cohort studies and a populationbased case-control study found a significant association between SHS exposure and the onset of asthma in adults. However, the authors concluded that because the evidence is scant and has potential problems in study design, “a definitive conclusion cannot be made at this time.” Subsequently, a population-based case-control study by Jaakkola et al. (2003) investigated the association between SHS exposure and the onset of adult asthma in the Pirkanmaa district of Finland. They recruited all incident cases of asthma in the district and selected populationbased controls. After excluding current or previous smokers, there were 239 lifetime nonsmoking cases and a comparison group of 487 lifetime nonsmoking controls. They found a twofold increased risk of asthma among those exposed to SHS in the home and the workplace compared with those who were unexposed. In a study published in 2006, Menzies et al. (2006) reported findings of a beneficial effect of the ban on smoking in Scotland on the health of bar workers. The investigators measured FEV1 level in nonasthmatic and asthmatic nonsmoking bar workers before and after the ban was implemented. They found that the FEV1 level increased after the ban was implemented.


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They also found a decrease in the prevalence of respiratory symptoms among nonsmoking bar workers. 19.5.2

Lung Function in Adults

With regard to involuntary smoking and lung function in adults, exposure to secondhand smoke has been associated in cross-sectional investigations with reduction of the FEF25–75. White and Froeb (1980) compared spirometric test results in middle-aged nonsmokers with at least 20 years of involuntary smoking in the workplace to the results in an unexposed control group of nonsmokers. The mean FEF25–75 of the exposed group was significantly reduced, by 15% of predicted value in women and by 13% in men. This investigation has been intensely criticized with regard to the spirometric test procedures, the determination and classification of exposures, and the handling of former smokers in the analyses. An investigation in France examined the effect of marriage to a smoker in over 7800 adults in seven cities (Kauffmann et al., 1983). The study included 849 male and 826 female nonsmokers exposed to tobacco smoke by their spouses’ smoking. At age above 40 years, the FEF25–75 was reduced in nonsmoking men and women with a smoking spouse. The investigators interpreted this finding as representing a cumulative adverse effect of marriage to a smoker. In a subsequent report, the original findings in the French women were confirmed, but a parallel analysis in a large population of U.S. women did not show effects of involuntary smoking on lung function (Spengler and Ferris, 1985). The results of an investigation of 163 nonsmoking women in the Netherlands also suggested adverse effects of tobacco smoke exposure in the home on lung function (Brunekreef et al., 1985; Remijn et al., 1985). Cross-sectional analysis of spirometric data collected in 1982 demonstrated adverse effects of tobacco smoke exposure in the home, but in a sample of women, domestic exposure to tobacco smoke was not associated with longitudinal decline of lung function during the period 1965–1982. Svendsen et al. (1987) assessed the effects of spouse smoking on 1400 nonsmoking male participants in the Multiple Risk Factor Intervention Trial (MRFIT). The subjects were aged 35–57 years at enrollment and were at high risk for mortality from coronary artery disease. At the baseline visit, the maximum FEV1 was approximately 3% lower for the men married to a smoker. Masi et al. (1988) evaluated lung function of 293 young adults, using spirometry and measurement of the diffusing capacity and lung volumes. The results varied with gender. In men, reduction of the maximal midexpiratory flow rate was associated with maternal smoking and exposure to SHS during childhood. In women, reduction of the diffusing capacity was associated with exposure to SHS at work. In the study of a general population sample in western Scotland, nonsmokers living with another household member who was a smoker had significantly reduced lung function in comparison to unexposed nonsmokers (Hole et al., 1989); the reduction of FEV1 associated with involuntary smoking was about 5%. Secondhand smokers with higher exposure had greater reduction of FEV1. Masjedi et al. (1990) investigated the effects of exposure to SHS on lung function of 288 nonsmoking volunteers living in Tehran. Ventilatory function was reduced significantly for men exposed at work, although an additional effect of exposure at home was not found. SHS exposure at home and at work did not reduce the lung function of the female subjects.

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In a meta-analysis of 15 cross-sectional studies, Carey et al. found a mean deficit of 1.7% in FEV1 level due to SHS exposure (Carey et al., 1999). They also conducted a separate crosssectional investigation of 1623 adults in Britain and found similar results, with a stronger effect in men than in women. Subsequently, Chen et al. found an inverse dose–response relationship between SHS and FEV1 level in 301 adults in Scotland for exposure to SHS at work (Chen et al., 2001). In a study of young Canadian adults, Jaakkola et al. (1995) did not find effects of home and workplace exposures on an 8-year change in lung function. In persons less than 26 years of age at enrollment, workplace SHS exposure was associated with greater decline. In another cohort study of 1391 lifetime nonsmokers and former smokers in California, Abbey et al. found a decrease in the ratio of FEV1 to FVC in both women exposed at home and men exposed at work (Abbey et al., 1998). However, the results were nonsignificant. Using NHANES III data, Eisner (2002) conducted a cross-sectional study to investigate the association between the level of SHS exposure and the pulmonary function in 10,581 nonsmoking adults and 440 nonsmoking adults with asthma. He found that FVC and FEV1 levels were significantly lower in adult females with the highest concentration of serum cotinine levels compared to adult females with lower serum cotinine levels. He did not find a significant association in adult males. Several investigators have reported associations of involuntary smoking with COPD in nonsmokers. In the Japanese cohort study, a nonsignificant trend of increasing mortality from chronic bronchitis and emphysema with increasing passive exposure of nonsmoking women was reported (Hirayama, 1984). Kalandidi et al. (1987) conducted a case-control study of involuntary smoking and chronic obstructive pulmonary disease; the cases were nonsmoking women with obstruction and reduction of the FEV1 by at least 20%. Smoking by the husband was associated with a doubling of risk. Dayal et al., (1994) conducted a casecontrol study of self-reported obstructive lung disease in 219 never-smoking residents of Philadelphia. Household SHS exposure from one or more packs per day was associated with a doubling of risk. In a prospective cohort study of 3914 nonsmoking Adventists, SHS exposure was associated with report of symptoms considered to be reflective of “airway obstructive disease” (Robbins et al., 1993). An association of SHS exposure with COPD seems biologically implausible, however, since only a minority of active smokers develop this disease, and adverse effects of involuntary smoking on lung function in adults have not been observed consistently (U.S. Department of Health and Human Services, 1984). The autopsy study of Trichopoulos et al. (1992) did show, however, that airways of nonsmokers can be affected by SHS. The 2006 report of the Surgeon General states that the evidence is suggestive but not sufficient to infer a causal association for the effects of SHS exposure on lung function in adults (U.S. Department of Health and Human Services, 2006). However, further research is warranted because of widespread exposure in workplaces and homes. 19.5.3

Odor and Irritation

Tobacco smoke contains numerous irritants, including particulate matter and gases (U.S. Department of Health and Human Services, 1986). Both questionnaire surveys and laboratory studies involving exposure to SHS have shown annoyance and irritation of the eyes and upper and lower airways from involuntary smoking. In several surveys of nonsmokers, complaints about tobacco smoke at work and in public places were common


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(U.S. Department of Health and Human Services, 1986): about 50% of respondents complained about tobacco smoke at work, and a majority were disturbed by tobacco smoke in restaurants. The experimental studies show that the rate of eye blinking is increased by SHS, as are complaints of nose and throat irritation (U.S. Department of Health and Human Services, 1986). In the study of exposure to secondhand smoke on commercial airline flights reported by Mattson et al. (1989), changes in nose and eye symptoms were associated with nicotine exposure. The odor and irritation associated with SHS merit special consideration because a high proportion of nonsmokers are annoyed by exposure to SHS, and control of concentrations in indoor air poses difficult problems in the management of heating, ventilating, and air-conditioning systems. Using a challenge protocol, Bascom et al. (1991) showed that persons characterizing themselves as SHS sensitive have greater responses on exposure than persons considering themselves as nonsensitive. They found similar results in a subsequent study (Bascom et al., 1996). Nowak et al. also found an increase in nose and mouth symptoms after exposing 10 persons with mild asthma to SHS in a chamber (Nowak et al., 1997). In a cross-sectional survey, Cummings et al. found a significant association between SHS exposure and irritation in 723 adult volunteers (Cummings et al., 1991). In 2001, Junker et al. reported findings of an experimental exposure assessment of SHS to determineodorandirritationthreshold levels(Junkeretal.,2001).They exposed people tovery low concentrations of sidestream SHS and measured odor detection, acute sensory symptoms, breathing patterns, and annoyance. They found that thresholds for odor and irritation were significantly lower than previously reported by 100 times and 10 times, respectively. 19.5.4

Total Mortality

Several cohort studies provide information on involuntary smoking and mortality from all causes. In the Scottish cohort study, total mortality was initially reported as increased for women living with a smoker, but not for men (Gillis et al., 1984). On further follow-up, allcause mortality was increased in all secondhand smokers (RR: 1.27; 95% CI: 0.95,1.70). As described previously, total mortality was also increased among nonsmoking participants in MRFIT who lived with smokers (Svendsen et al., 1987). In contrast, mortality was not increased for nonsmoking female subjects in a study in Amsterdam (Vandenbroucke et al., 1984). Neither the study in Scotland nor the study in Amsterdam controlled for other factors that influence total mortality. In the cohort study in Washington County, all-cause mortality rates were significantly increased for men (RR: 1.17) and for women (RR: 1.15) after adjustment for housing quality, schooling, and marital status (Sandler et al., 1989a). Allcause mortality was also increased for secondhand smokers in the Evans County cohort (RR: 1.39, 95% CI: 0.99,1.94). Wells (1988) has made an estimate of the number of adult deaths in the United States attributable to secondhand smoke exposure. The total is about 46,000, including 3000 from lung cancer, 11,000 from other cancers, and 32,000 from heart disease. Additionally, Nurminen and Jaakkola estimated adult mortality attributable to SHS exposures in the workplace in Finland (Nurminen and Jaakkola, 2001). They estimated that 250 deaths were due to exposure to secondhand smoke at work in 1996. Among these deaths, 2.8% were from lung cancer, 1.1% from chronic obstructive pulmonary disease, 4.5% from asthma, 3.4% for ischemic heart disease, and 9.4% for cerebrovascular stroke. The small excesses of all-cause mortality associated with exposure to secondhand smoke in the epidemiological studies parallel the findings for cardiovascular disease, the leading

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cause of death in these cohorts. The increased risk of death associated with secondhand smoke has public health significance as an indicator of the overall impact of this avoidable exposure. 19.5.5

Control Measures

Since the 1980s, there has been growing momentum for making public places and workplaces smoke free. The public health basis for this movement lies in the increasingly strong findings on the health risks of SHS (Table 19.1) and on the need to eliminate smoking indoors to fully protect nonsmokers from inhaling SHS. Cigarettes are strong sources of gaseous and particulate emissions, and use of mass balance models implies that concentrations of SHS components could not be controlled by either ventilation or air cleaning (U.S. Department of Health and Human Services, 2006). Such consideration led the American Society of Heating, Refrigerating and Air Conditioning Engineers (ASHRAE) to conclude that ventilation was not a sufficient control measure for SHS (Samet et al., 2005). The 2006 report of the Surgeon General reached a similar conclusion (U.S. Department of Health and Human Services, 2006). There are regulatory and nonregulatory approaches to eliminating smoking indoors. There are an increasing number of local and state ordinances banning smoking in public places and workplaces, and many large companies have policies in place that prohibit smoking indoors. Even some major hotel chains are now smoke free. In recent surveys, the majority of employed people report working under a smoke-free policy, although there is some variation by type of workplace and region of the country (U.S. Department of Health and Human Services, 2006). Blue collar and farm workers are least likely to be covered. The home is not subject to regulation, but increasing numbers of households in the United States have voluntary policies in place (U.S. Department of Health and Human Services, 2006). The majority of households are now smoke free, and increasing numbers of households with smokers have smoke-free policies in place. The effectiveness of these policies is still inadequate, and research is in progress to develop more efficacious approaches to reduce SHS exposure in homes, particularly, for example, for children with asthma who are especially susceptible to the adverse health effects of SHS. Thischaptersummarizestheconvergingandnow extensiveevidenceonthehealtheffectsof involuntary exposure to tobacco smoke. Although the initial research on involuntary smoking addressed respiratory effects, subsequent investigations have examined associations with diverse health effects including nonrespiratory cancers in children and adults, ischemic heart disease, age at menopause, sudden infant death syndrome, and birth weight. The evidence on involuntary exposure to tobacco smoke is now voluminous and consequently this review is selective in its citations. The most recent compilation of the evidence can be found in the 2005 report of the California Environmental Protection Agency, “Health Effects of Exposure to Environmental Tobacco Smoke” (California Environmental Protection Agency and Air Resources Board, 2005) and the 2006 report of the U.S. Surgeon General (U.S. Department of Health and Human Services, 2006).

19.6 SUMMARY The effects of active smoking and the toxicology of cigarette smoking have been comprehensively examined. The periodic reports of the U.S. Surgeon General and other summary

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reports have considered the extensive evidence on active smoking; these reports have provided definitive conclusions concerning the adverse effects of active smoking, which have prompted public policies and scientific research directed at prevention and cessation and smoking. Although the evidence on involuntary smoking is not as extensive as that on active smoking, health risks of involuntary smoking have been identified and causal conclusions reached, beginning in the mid-1980s (Table 19.1). The 1986 report of the U.S. Surgeon General (U.S. Department of Health and Human Services, 1986) and the 1986 report of the National Research Council (National Research Council and Committee on Passive Smoking, 1986) both concluded that involuntary exposure to tobacco smoke causes respiratory infections in children, increases the prevalence of respiratory symptoms in children, reduces the rate of functional growth as the lung matures, and causes lung cancer in nonsmokers. These conclusions have been reaffirmed in subsequent reports (California Environmental Protection Agency and Office of Environmental Health Hazard Assessment, 1997; U.S. Department of Health and Human Services, 2006; U.S. Environmental Protection Agency, 1992) and new conclusions added. Involuntary smoking is now considered as a cause of asthma and a factor exacerbating asthma (California Environmental Protection Agency and Office of Environmental Health Hazard Assessment, 1997; U.S. Department of Health and Human Services, 2006; U.S. Environmental Protection Agency, 1992) and as a cause of heart disease (California Environmental Protection Agency and Office of Environmental Health Hazard Assessment, 1997; U.S. Department of Health and Human Services, 2006). The 2006 Surgeon General’s report (U.S. Department of Health and Human Services, 2006) leaves no doubt: secondhand smoke causes premature death and disease in children and adults who do not smoke (p. 11). The adverse effects of involuntary exposure to tobacco smoke have provided a strong rationale for policies directed at reducing and eliminating exposure of nonsmokers to SHS (U.S. Department of Health and Human Services, 1986). Complete protection of nonsmokers in public locations and the workplace may require the banning of smoking, since the 1986 report of the Surgeon General (U.S. Department of Health and Human Services, 1986) concluded that “the simple separation of smokers and nonsmokers within the same air space may reduce, but does not eliminate, the exposure of nonsmokers to environmental tobacco smoke” (p. 7). The 2006 Surgeon General’s report went even further: “the scientific evidence indicates that there is no risk-free level of exposure to secondhand smoke” (p. 11).

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20 LEAD AND COMPOUNDS Lester D. Grant*

20.1 INTRODUCTION Lead (Pb), a heavy metal with numerous useful properties (low melting point, highly malleable, etc.), has been put to many diverse uses by both ancient civilizations (e.g., the Roman Empire) and by modern societies. The extensive expanded use of the metal in modern times (in water distribution systems, in additives to paints and gasoline, for electronics applications, etc.) caused widespread increases in lead exposures for human populations around the world, especially during the twentieth century. Despite its broad usefulness, however, the metal has also been long recognized to be acutely toxic at high-dose exposure (e.g., see Aub et al., 1925; Beechx, 1986). Diagnosis of classically defined acute lead poisoning (often life-threatening) historically typically involved clinical observation in individual medical cases of signs and symptoms of (a) marked impairment of red blood cell formation/function; (b) severe central and/or peripheral nervous system (PNS) damage/functional impairment; and/or (c) notable kidney damage/renal dysfunction. Very importantly, however, extensive new research findings emerging since the 1970s (comprising major advances in our understanding of lead toxicity) indicate that lead can exert toxic effects of concern on many organ systems and at exposure levels far lower than those producing clinically evident signs and symptoms of overt lead intoxication. Such “subclinical” lead toxicity effects are not identifiable through routine clinical examination of any one patient at one point in time, but rather their characterization generally requires more sophisticated methods, for example, identification through longitudinal observation. Also, although many such subclinical effects may be very subtle, their ultimate impact on a population basis can be substantial, as illustrated both for lead (Eckerman et al., 1999) and for other chemicals (Lester et al., 1998). *

Formerly Director (now retired), National Center for Environmental Assessment–Research Triangle Park Division (NCEA-RTP), U.S. Environmental Protection Agency (U.S. EPA). The contents of this chapter have not undergone review or clearance by U.S. EPA and should not be taken to represent views or official positions of the U.S. EPA.


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Demonstration during the past 30–40 years of “subclinical” toxic effects at lower and lower exposure levels (and recognition of their potential public health implications) has led to repeated downward revision over the same time period of views regarding lead exposure levels seen as posing unacceptable risks for adverse human health impacts. Stimulated by such changes in viewpoints regarding “safe” lead exposure levels, lead regulatory guidelines/standards have been successively tightened and extensive remedial actions implemented in many parts of the world, producing notable reductions in lead exposure among many population groups. Thus, although clinically evident acute lead intoxication due to high-dose exposures continues to be of much concern in certain geographic areas, subclinical lead toxicity has come to be of much broader public health interest. Accordingly, greater attention is devoted here to lower level lead exposure effects. Also discussed here are factors affecting susceptibility/vulnerability to such effects, some hypotheses and evidence regarding potential mechanisms of action, and progress in developing biokinetic models used to predict likely increased risk of lead-related toxicity among human population groups due to various lead exposure scenarios. The remarkable advances made in our knowledge of lead toxicity during the past 30–40 years have been described in numerous reviews published during the last several decades. The prior edition of this chapter, for example, noted those by Davis and Svendzgaard (1987) and Lippmann (1990), and a number of newer, more recent reviews are cited at various points later in this chapter. Major advances made via evolving lead research during the past 4–5 decades are especially well reflected by the series of intensively peer-reviewed, periodic assessments of the latest available lead-related information that comprises U.S. Environmental Protection Agency (U.S. EPA) Air Quality Criteria for Lead documents and associated Addendum/Supplement materials (U.S. EPA, 1977, 1986a, 1986b, 1990a, 2006) generated to support the setting and periodic review of U.S. National Ambient Air Quality Standards (NAAQS) for Lead. The reader is referred to such EPA documents and other reviews for more extensive, detailed discussions of particular topics. It should also be noted that certain important points made in older classic literature, as well as advances made in our knowledge of lead toxicology through the late 1990s/early 2000s, were delineated well in the previous edition of this chapter (by Mahaffey, McKinney, and Reigart), the contents of which still largely remain germane today. Extensive portions of text from that edition are, therefore, retained here largely intact or with minimal revision. Other text has been updated to reflect more recent findings on a given topic or to provide information on additional topics beyond those covered in the prior edition, with the most recent EPA assessment (U.S. EPA, 2006) of lead exposure and toxicity information being drawn upon as a particularly cogent, readily accessible authoritative source.

20.2 PHYSICAL/CHEMICAL PROPERTIES AND BEHAVIOR OF LEAD AND ITS COMPOUNDS Lead is a member of subgroup IVA of the periodic table and is a typical heavy metal (Greninger et al., 1978) with a relatively high atomic weight. The valence shell of the lead atom in the ground state has two s and two p electrons. Because the element has four electrons in its outer shell, lead would be expected to show a normal valence þ4 in its compounds. The two s electrons of lead do not readily ionize and are thus often referred to as the inert pair. Lead is considered to have a stable oxidation state (Pb2þ) that furnishes a divalent ion. The metabolism of lead in a redox sense does not appear to be particularly important

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in determining its biological properties. Bivalent lead has a notable tendency to form well-delineated, often highly crystalline basic salts of both anhydrous and hydrated types, for example, white lead (a pigment of past widespread commercial use). In general, the inorganic chemistry of bivalent lead resembles that of alkaline earth elements. Several lead salts (e.g., lead carbonate, nitrate, and sulfate) are isomorphous with corresponding strontium and barium compounds. Lead forms highly insoluble salts of phosphate, carbonate, and sulfide. Lead can also form salts with organic acids, which is the basis for the use of certain chelating agents for treating lead intoxication. Elemental lead would be readily oxidized in biological systems and is thus not considered as a separate form here. Lead’s position in the periodic chart favors formation of covalent rather than ionic bonds in Pb4þ compounds. This expectation is confirmed by the properties of compounds such as lead tetrachloride and tetraacetate. Predominately covalent bonding is also seen with organolead compounds (with up to four Pb C bonds). An organometallic compound differs fundamentally in both chemical and biological properties from an ionic compound of the same metal. Thus, determining only the total amount of the metal in a biological sample can be very misleading with regard to estimating the potential for toxic effects. In general, inorganic lead has been more extensively studied than organometallic lead. Speciation and movement of lead and other heavy metals in the environment were extensively discussed at a mid-1980s international conference (Landner, 1987), and the biological effects of organolead compounds were reviewed by Grandjean and Grandjean (1984). An appreciation of some general features of metal chemistry is very helpful in understanding specific aspects of lead chemistry (Hanzlik, 1981). Among the most important criteria differentiating metal ions from each other, and from electrophilic organic species, is the chemistry of their bonding to biological ligands. Metal–ligand bonds can be as strong, in a thermodynamic sense, as bonds formed when a reactive epoxide alkylates a nucleophilic group in DNA or protein. Regardless of the mechanism by which a metal ion enters a biological system, complexation undoubtedly plays a role in both its distribution within and elimination from the organism. Metals’ ions are Lewis acids, and one very important determinant of their affinity for ligands is their charge/radius ratios. Increasing the metal’s oxidation state increases its Lewis acidity and its affinity for a given ligand (assuming that it does not ionize the ligand). In an antagonist sense, the relative size of the ion can also be important. In addition to the energetics of the ligation of ions, the energetics of the ionization process itself must also be considered. In this case, complex geometry and ligand exchange rate play important roles. Most metal complexes undergo ligand exchange by processes that involve a dissociative rate-limiting step analogous to the Sn1 solvolysis of alkyl halides. For a given metal, the rates are rather sensitive to the nature of the departing ligand and are essentially independent of the entering group. The dependence of biological activity on ligand exchange rates reflects the fact that the complex must be sufficiently inert to survive long enough in vivo to reach critical reactive target molecules and yet, once having reached those sites, be sufficiently labile to react. A given metal-macromolecular interaction may persist enough to have biological consequences, if the equilibrium constant of the rate for dissociation of the complex is very small. The hard–soft, acid–base dichotomy provides a rationale underlying many features of the behavior of metal systems in chemistry and biology. This parameter is correlated qualitatively with the charge–size ratio of the ion in that large ions of low ionic charge have easily polarizable or deformable (i.e., soft) electrostatic fields about them, whereas small highly

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charged ions with relatively intense electrostatic fields are hard. Flexibility in hard–soft ligand preferences appears to be a key property underlying the biological activity/toxicity of metals in living systems. Several toxic heavy metal ions (i.e., Pb2þ, Hg2þ, Th1þ, Ni2þ, and Sb3þ) are classified as soft or have borderline properties in this classification scheme. The ligands, analogously, can also be classified by these hard–soft criteria. In this regard, it is not surprising that the rate of transport of a given metal ion across model liquid membranes can be varied by several orders of magnitude simply by altering the anion (ligand) present in the original salt solution (Christensen et al., 1978). Like the other group IVA metals, tin and germanium, lead forms complexes in which the donor atom is chiefly oxygen (Greninger et al., 1978). It also forms stable complexes with sulfur and halogens as the donor. Carbon and nitrogen donors are less common. Lead generally forms complexes with the coordination number six (having octahedral geometrical structure), whereas other geometries are less common. On the basis that hard metal ions prefer to bind ligands and vice versa, one might expect lead to bind halides in the order I 4 Br 4 Cl 4 F. This is consistent with the early use of potassium iodide to enhance lead removal from the body (Aub et al., 1925). Another important aspect of metal chemistry is the potential for metal compounds to act as initiators or catalysts in vivo (Hanzlik, 1981). It is not difficult to envision that inhibition of enzyme molecules by stoichiometric quantities of tightly bound metal ions could reduce the flow of vital metabolites through a pathway and, thus, cause toxicity. In addition to stimulating or inhibiting the synthesis of enzymes, as well as the enzymes’ activities, many simple metal ions and compounds have catalytic activity in their own right. Of much importance here are the electrochemical gradients across biological membranes and the potential of a foreign metal ion to act as an “antimetabolite.” This may be significant in view of possible existence of a mechanism for coupling biological oxidation–reduction pathways to ion transport and the control of membrane potential. In many cases, apparently nonessential metals are absorbed into an organism and not excreted at all; rather, they are simply concentrated and deposited in granular, insoluble complexes with or without accompanying proteinaceous material. There are several ways to express the relationship of lead to other metals, both foreign and endogenous. For example, the resemblance of bivalent lead chemistry to that of the alkaline earth metals in general was mentioned earlier. Of particular note is the similarity to calcium. Lead and calcium both form insoluble carbonates and insoluble phosphates. However, lead phosphate is much more insoluble than lead carbonate, whereas calcium phosphate is more soluble than calcium carbonate. Lead phosphate is one of the few insoluble phosphates that do not react with most chemical reagents. The extreme insolubility of lead phosphate may serve as a driving force for lead to function as a phosphate scavenger in biological systems, which may include inorganic forms of phosphate as well as the various important phosphate esters. The ultimate deposition of lead in the skeleton is consistent with lead’s chemical relationship to calcium and the formation of highly insoluble salts (Aub et al., 1925). Strontium, another alkaline earth metal, can also compete with calcium in bone tissue (Smith et al., 1985), and this chemical similarity of lead and strontium may be related to their ionic radii and the stable 2þ oxidation state relative to that of Ca2þ. Intestinal calcium-binding proteins have been shown to bind lead with high affinities and in preference to calcium (Fullmer et al., 1985). These proteins bind several other cations (notably Sr2þ, Ba2þ, and Cd2þ) in a fashion apparently related to metal ionic radii relative to calcium. Another important factor that can determine metal chemistry and biology is the hard–soft, acid–base property (amphoteric nature). In this regard, lead is similar to iron, copper, zinc,

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mercury, and thallium, among others. The amphoteric property, along with redox cycling, appears to account, in large part, for the importance of iron in biological systems and its adsorption, storage, and transfer in these systems (Hanzlik, 1981). Such effects are attributable to the greater polarizabilities (function of number of electrons) of these cations and subsequently greater covalent character of the bonds they form with donor ligands relative to that expected for alkaline earth cations of the same size. The situation with Pb2þ is very analogous to that of Th1þ (Izatt et al., 1976). Similarities in size and coordination chemistry may be important factors that determine the ability of metals to act antagonistically (Hill and Matrone, 1970). The chemical similarity of lead to certain alkaline earth metals (particularly calcium) and lead’s ability to form highly insoluble salts (particularly of carbonate and phosphate) along with increased affinities to biological donors (enriched in oxygen and possibly nitrogen) due to favorable polarizabilities may well account for much of the relevant biological/toxicological chemistry of lead compounds. In an overall sense, the importance of the ligand exchange chemistry of divalent lead is emphasized in the expression of toxicity. Ligands can include simple anions or more complex donors that can form chelates or organic complexes. The biological activity of a given metal is a consequence of the way in which the metal’s compounds (salts, complexes, etc.) and cells interact. It is this interaction that is governed by intrinsic chemical properties (modulated by certain physical properties) of both the particular metal compound and the cell. In addition, the extent of cellular interaction can be affected by the same or different chemical properties that determine the in vivo absorption, distribution, and elimination of the compounds. Importantly, lead forms highly stable bonds with sulfur and sulfur-containing compounds, but somewhat less stable ones with carboxylic acids (O-based ligands) and imidazoles (N-based ligands) (Claudio et al., 2003), as noted in U.S. EPA (2006). Also, as noted there, lead competes very effectively in biological systems with native or homeostatic metal ions for binding with sulfhydral, carboxyl, and imidazole side chains that comprise enzyme active sites, and this competition leads to inhibition of enzyme activity, the replacement of calcium in bone, and many other deleterious health impacts. Other key features of lead coordination chemistry, their roles in biological systems, and relationships to lead-induced adverse health effects are delineated in the review by Claudio et al. (2003).

20.3 LEAD IN THE ENVIRONMENT AND HUMAN EXPOSURE The previous edition of this chapter noted that trace metals, such as lead, can be present in the environment in various forms (Boline, 1981), such as free hydrated ions; ion pair salts/ complexes; organic complexes/chelates; surface-adsorbed material; and undissolved compound. Although differences may occur in valence states and associated ligands (including mixed ligands) across these various forms, the metal identity is still retained, and the chemical reactivity of the metal is a function of the combined physicochemical properties of the metal and its associated ligands. This chemical reactivity can be modulated by physical properties, such as the surface properties of the metal compound itself. Thus, under certain physical, chemical, and biological conditions, it is possible for a given metal to assume more than one form, which can follow new pathways of chemical reactivity. This same reactive potential contributes to the posing by certain metals of possible toxic threats to the environment and living systems that concentrate them. The range of chemical properties and reactivities associated with various types of metal compounds is thus

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much greater than that of simple organic compounds. Since the metal is never destroyed, the potential to exert this complex chemistry always exists. Inorganic lead compounds to which humans are likely to be exposed include halides and oxides, sulfides and sulfate, carbonate, and chromate. With the exception of a few sporadic measurements in air, marine fish, sediments, birds, and in human brains, there is relatively little information available on organic lead compounds (Jawrerski et al., 1987). It appears that most organic lead compounds in the environment come from the release of organolead compounds (such as those used as gasoline additives) prior to or during their use rather than being derived from inorganic compounds of lead. In the expanding number of areas restricting the use of lead-based gasoline additives in the past 20–30 years, an increasingly greater proportion of inorganic lead compounds have come to dominate. Thus, the global movement of inorganic lead and its compounds, as well as human exposure to them, are of much greater concern, have been far more extensively studied, and are the main focus here. Calculations of loading rates of lead and other trace metals into various environmental compartments indicate that human activities exert major impacts on global and regional cycles of most trace elements (Nriagu and Pacyna, 1988). Other environmental problems, such as acidic precipitation (Mohnen, 1988), have contributed to mobilization and distribution of metals in the environment as well. The greatly increased circulation of toxic metals through the air, soil, and water, and their ultimate transfer into the human food chain remain important environmental issues and likely entail unknown health risks for future generations. The biogeochemical cycling of lead and routes of human exposure were described earlier by Schlag (1987) and have been more recently extensively discussed in U.S. EPA (2006). Key types of environmental reservoirs of lead can be identified and quantitative estimations made of inputs to them from various natural or human sources in a reasonably straightforward manner, but rates of transfer within and between the reservoirs are generally known only qualitatively or semiquantitatively. For example, as noted in the prior edition of this chapter, various estimates of natural and anthropogenic lead emissions to air and to the oceans have been made on a global basis (e.g., by Nriagu and Pacyna, 1988), and all such estimates indicate that contributions from anthropogenic sources are at least one to two orders of magnitude greater than from the natural ones. Detectable long-term elevations in global lead emissions can be seen starting as of the time of the Roman Empire, followed by a relatively slow increase over many centuries and then a steep rise starting in the eighteenth century, which peaked at around 400,000 tons per year during 1970–1980 and has since declined to about 100,000 tons per year (Nriagu, 1998). This temporal pattern of changing worldwide lead emissions is reflected well by examination of natural historical records of lead accumulated over time in ice packs or other layered natural materials, with analogous peaking in the 1970–1980 period being followed by declining lead levels. For example, the pattern of atmospheric lead deposition in a peat bog in the Swiss Jura Mountains parallels well historical increases in deposition of lead attributable to the introduction in 1947 of leaded gasoline into Switzerland, with later bog layer samples corresponding to 1991 showing a decline in atmospheric lead deposition and a shift in isotopic ratios toward radiogenic values (Shotyk et al., 1998) reflective of subsequent phasing out of leaded gasoline in Europe. Analogous patterns of variations in sediment lead levels (i.e., reaching peak concentrations in layers reflecting highest deposition during the 1970s, followed by ensuing declines) have been observed for a variety of other locations in North America and Europe. For more detailed reviews of natural historical records reflecting temporal


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variations in lead emissions and deposition, see Boutron et al. (1994), Weiss et al. (1999), and Garty (2001). Human lead exposure generally occurs via one or more of four main components of the human environment: inhaled air, soil and dust of various types, drinking water, and food. The prior edition of this chapter noted that the primary medium for widespread dispersal of lead in the ambient (outdoor) environment tends to be air, because lead-containing fine particles (emitted mainly by high-temperature anthropogenic sources) can travel long distances before settling out via wet, dry, or cloud deposition. Lead deposition from air is most intense near a given source, but the zone of readily detectable elevated deposition can extend some distance (even many kilometers) away. Substantial decreases in ambient airborne lead levels have occurred in parallel with the phase down in usage of organic lead compounds as additives in gasoline in many countries, for example, more than 90% decrease in U.S. ambient air lead levels from the mid-1970s to more current urban air concentrations typically falling in the 0.10– 0.25 mg/m3 range by 2000–2002, as noted in U.S. EPA (2006). However, accumulation of lead in roadside soil and other soils due to past deposition of airborne lead from gasoline, smelters, and other sources constitutes a persisting reservoir of anthropogenically generated lead that will likely continue for many decades to contribute to human exposure and health risks. This is partly due to the fact that most lead particles deposited on soil are retained and, eventually, are mixed into the surface layer, with the lead accumulated at the soil surface becoming available to be taken up by plants, grazing animals, or soil microorganisms and, thereby, enter terrestrial food chains. Also, very importantly, direct exposure of children to lead in soil can occur by oral intake of lead-contaminated dust and dirt during normal hand-to-mouth activity. In addition, lead-contaminated soil particles can be resuspended in air and as such become a long-term source of airborne lead exposure for humans via the inhalation route as well. Lead in rivers comes from runoff, erosion, and direct deposition from air. Freshwater generally contains more inorganic and organic suspended particulate material than marine water, and this suspended material has a strong tendency to adsorb any dissolved lead, with lead adsorption potential being much greater for smaller size particles (due to their notably greater surface area per unit weight of lead) than for larger particles (Rhoads and Cahill, 1999). Lead adsorption into sediments can be enhanced by the presence of various substances. For example, lead concentrations in sediment typically increase with humic (organic matter) content (Kiratli and Ergin, 1996; Rhoads and Cahill, 1999), and lead in sediments can be sequestered on iron or manganese oxides (Schintu et al., 1991; Peltier et al., 2003; Gallon et al., 2004) or its adsorption increased by sulfides, especially under anoxic conditions (Kiratli and Ergin, 1996; Perkins et al., 2000). Most of the lead entering the open oceans comes from atmospheric deposition rather than from rivers. The lower concentrations of particulate matter in marine waters, as well as high salt (chloride, bromide, etc.) levels, tend to favor a larger proportion of dissolved lead in the marine water column. Deep ocean sediments may thereby represent a sink for lead, since there is little evidence to suggest notable remobilization from such sediments. On the other hand, it appears that lead in sediments in freshwater lakes and rivers can be remobilized (e.g., see Steding et al., 2000; Hlavay et al., 2001; Peltier et al., 2003; Kurkjian et al., 2004; Gallon et al., 2004), especially under acidic conditions. Thus, in some areas, lead in freshwater sediments may continue to be a potential source of water-related exposure to lead deposited many years earlier from the air, via runoff, and so on. Lead in drinking water supplied by municipal water distribution systems typically derives mainly from corrosion of lead pipes, lead-based solder, or bronze and brass fixtures

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(e.g., faucets) used as part of plumbing within residences or workplaces, with little lead generally coming from properly buffered utility supplies (Lee et al., 1989; Singley, 1994; Gulson et al., 1994; Isaac et al., 1997). Low pH (acidic) water enhances the leaching of lead from indoor plumbing components, which can be greatly reduced by buffering utility water supplies toward neutral pH conditions. Of much importance, the addition for disinfection purposes of chlorine to drinking water (which can lower the pH) does not generally increase leaching of lead into the water, because of a chlorine reaction with Pb2þ that results in precipitating out of a highly insoluble, red-brown colored lead solid (Edwards and Dudi, 2004). On the other hand, the insoluble lead solid is not formed in the presence of chloramines, and the introduction of chloramine disinfectants in place of chlorine (especially in the absence of other adequate water chemistry adjustments) appears to have contributed to increases in drinking water lead levels seen in some U.S. communities since 2000, as noted in U.S. EPA (2006). Potential human exposure via drinking of lead-contaminated water has received much attention in the past and can still be a nontrivial contributor to lead exposures for some populations. As an example, tap water lead was the main correlate of elevations in maternal blood lead levels in a study of mothers and infants in Glasgow, Scotland (Watt et al., 1996). In a U.S. prospective study, Lanphear et al. (2002) found that children exposed to water with lead levels over 5 ppb had blood lead levels 1.0 mg/dL higher than children exposed to water with lead concentrations less than 5 ppb. Although exposure to lead via drinking water undoubtedly still occurs, it is often difficult to readily determine the importance of this exposure route in contributing to any specific overall toxic insult, due in part to the considerable potential for wide variations in the concentration and bioavailability of lead in water. Furthermore, the bioavailability of lead and other metals in water not only depends on trace metal solubility but also on numerous complex chemical equilibria affected by the presence of other trace inorganic and organic compounds in the water (Jackson and Sheiham, 1986). The bioavailability of lead in soils, food, and inhaled air may depend on similar factors that determine the ligand exchange chemistry once in contact with the biota and aqueous phase. Accordingly, metal speciation analyses and solubility modeling are likely to yield further insights and improved understanding of the potential for toxic insult (Hunt and Creasey, 1980) and, thereby, contribute to improved scientific bases for abatement strategies for lead and other metal pollutants (e.g., reducing the bioavailability of trace metals in water via manipulation of their solubility and aqueous chemistry, as suggested by French and Hunt, 1988). Lead in food derives from plant and animal exposure to contaminated air, soil, and water, and from products used in the processing and storage of foods. The prior edition of this chapter noted that lead in foods in many countries has declined dramatically during the last several decades, from typically 100 to 200 mg/day (Mahaffey, 1977) to typically less than 5 mg/day by the mid-1990s (Bolger et al., 1996). This decline resulted from virtual bans on the use of lead solders in food and beverage containers and to more widely practiced limits on the use of lead glazes in pottery and food storage containers. Lead contamination of raw food resulting from air and water contamination has also notably declined. Marked decreases noted above for lead concentrations in several environmental media reflect, in large part, strong steps taken to reduce environmental lead levels, based on the growing recognition that the best approach for dealing with lead intoxication as a public health issue is to prevent lead exposure. This notable progress in reducing lead in environment media during the last several decades represents a major success story for a number of countries. For example, the prior edition of this chapter highlighted dramatic reductions in


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blood lead levels among U.S. population groups. That is, between NHANES II (conducted 1976–1980) and Phase I of NHANES III (conducted 1988–1991), the geometric mean blood lead level for U.S. persons aged 1 through 74 years declined from 12.8 mg/dL to 2.9 mg/dL, and the prevalence of elevated blood lead levels (i.e., 410 mg/dL) decreased from 77.8% to 4.4% (Mahaffey et al., 1982; Pirkle et al., 1994). Also, between NHANES III Phase II (1988–1991) and Phase III (1991–1994), the comparable geometric mean decreased by 22% (Centers for Disease Control, 1997). It was further noted that the blood lead concentrations of children had been reduced by more than 80% over the prior 2 decades; and U.S. EPA (2006) more recently noted that marked decreases in environmental lead during the 1980s–1990s were paralleled by decreases in concurrent blood lead levels of U.S children (from a geometric mean of 15 mg/dL in 1980 to 1–2 mg/dL in 2004). Such declines in blood lead levels among the U.S. general population and children not only partly reflect changing economic circumstances (e.g., declining numbers of operating U.S. primary or secondary lead smelters) but also, very importantly, the success of highly effective primary interventions (virtual elimination of lead additives from gasoline, shifting to low lead plumbing solder and fixtures, removal of lead solder from food and beverage cans, etc.) and secondary prevention strategies, for example, public health screening/ education programs and improved nutrition. Despite this overall success, however, some U.S children still experience blood lead concentrations above10 mg/dL, and disproportionate numbers of black and low-income children continue to exhibit such elevated blood lead levels. The most extensive remaining source of lead exposure for U.S. children appears to be lead-based paint in housing, especially for those living in older residences, which are more likely to contain lead-based paints (Jacobs, 1995). Although vigorous national efforts to reduce lead exposure from leaded paint in housing still continue, the problem persists, with the CDC (1997) having estimated in the 1990s that 890,000 U.S. children had sufficiently high blood lead levels (10 mg/dL or above) to impair their learning ability. The prior edition of this chapter noted that elevated blood lead concentrations occur occasionally among adults, with the use of “folk” remedies, cosmetics, and lead-glazed pottery typically being reported to be the source in such isolated cases among the general population. Further, the remaining most persistent cause of elevated blood lead concentrations among adults was noted to be occupational exposures. Apropos to this, the National Institute for Occupational Safety and Health (NIOSH) was also noted as maintaining an Adult Blood Lead Epidemiology and Surveillance (ABLES) Program that tracks the laboratory-reported elevated blood lead levels among adults in the United States, and based on data reported from the 27 states, the cumulative number of reports in 1996 was 16,551 adults with blood lead concentrations of 25 mg/dL or higher, with 318 blood lead concentrations greater than or equal to 60 mg/dL (NIOSH, 1998). More recently, the ABLES Web site currently indicates that the geometric mean of all reported U.S adult blood lead concentrations is now less than 3 mg/dL, an average much lower than the 25 mg/dL level that the U.S. Department of Human Service recommends not to be exceeded by adults. Adverse health effects of lead obviously derive from various types and intensities of responses of cells following particular patterns and intensities of lead exposure. The transfer of lead from the environment to the cells and subsequent interactions of lead with cell components occur as functions of the physical/chemical properties of lead and of physiological factors inherent in the organism. The delineation of physiological pathways involved in determining uptake and internal distribution of lead to various tissues, lead’s accumulation in particular types of tissue and internal redistribution, as well as factors affecting uptake, internal distribution, metabolism, or excretion of the metal are all important,

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especially (a) to help identify tissues and mechanisms underlying different types of toxic effects of lead and (b) to enhance our ability to estimate or predict the likelihood that adverse responses will result from varying lead exposure patterns and intensities. The human interface with the environment that permits lead entry into the body is comprised of the gastrointestinal (GI) system, respiratory system (including the nasal cavity), and skin. The absorption of lead via these portals of entry and the important factors that affect absorption of lead from different exposure routes as well as certain other aspects of the biokinetics of lead are summarized next, as a prelude to discussion of lead-related health effects. More detailed reviews of lead toxicokinetics can be found in Mushak (1991, 1993, 1998) and U.S. EPA (1986a, 1986b, 2006).

20.4 LEAD ABSORPTION 20.4.1

Gastrointestinal Absorption of Lead

The mechanisms involved in the gastrointestinal absorption of lead are only partially understood. Most of the pertinent research on such mechanisms has been carried out in rodents, especially the rat. Findings by Aungst and Fung (1981), Henning and Cooper (1988), and Fullmer (1997) emphasize the complexity of the process and the extent to which the specific research findings depend on various factors, for example, the dose of lead and the physiological state, nutritional condition, and age of the exposed animal. Based on data from several types of experiments, at least two mechanisms for GI lead absorption appear to exist. One exhibits characteristics of energy-dependent, carriermediated, active transport (Aungst and Fung, 1981); as has been reported, whether this process is saturable or not depend at least in part on the dose of lead used (Keller and Doherty, 1980a, 1980b). This absorption mechanism has active transport mechanism characteristics, because intestinal uptake and flux of lead depend on metabolic energy. At a buffer lead concentration on the mucosal side of 0.5 mM, capacity-limited processes contributed nearly 200 times more to the mucosal-to-serosal lead flux than did diffusion (Aungst and Fung, 1981). At a lead concentration three orders of magnitude higher (48.3 mM), diffusion still only accounted for less than 20% of the flux (Aungst and Fung, 1981). Others have concluded that the major control level for GI absorption of lead likely resides in the intestinal mucosal cell and that the interrelationships between the elements may affect lead bioavailability at both luminal and mucosal levels. It has also been suggested that there are three components to the absorptive phase: uptake by the mucosal cell, transfer through the cell, and movement into the plasma (Ragan, 1983). Lead absorption also appears to have a dose-dependent component. In addition to the work of Aungst and Fung (1981), other work includes in situ studies by Barton et al. (1978a, 1978b) in which the percentage of lead absorbed depended on the magnitude of the dose and was increased in iron-deficient animals (Hamilton, 1978) in relation to fed or iron-replete animals. 20.4.2

Effects of Age on Lead Absorption

Age substantially influences absorption of lead in human and nonhuman primates. For example, Willes et al. (1977) reported that infant monkeys at 10 and 150 days of age retained 64.5% and 69.8% of an oral dose of 210 PbðNO3 Þ2 , whereas adult monkeys only retained 3.2% of an oral dose. Similar age-related differences have been observed for humans. Using

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classic balance study techniques, Kehoe (1961) found the GI absorption of lead by adult males to be 5–10% of ingested lead. This range of lead absorption with usual patterns of food intake (i.e., not fasting) has been confirmed to be in the range of 5–15% of ingested lead, based on studies with short-lived lead radioisotopes (Hursch and Suomela, 1968) or stable lead isotopes (Rabinowitz et al., 1975, 1976). There are few studies of adult female subjects. Only James et al. (1985), whose subjects (age 26–77 years) included both females and males (12 women and 11 men), reported lead absorption from foods and beverages. However, the report did not discuss any sex-related differences observed between the retention rates for radio-labeled lead. Oral exposure uptake rates for children have been much less clearly documented than those for adults. Available absorption coefficients for children have been derived mainly from two mass balance studies with small numbers of children. Alexander et al. (1974) conducted balance studies in eight subjects (aged 3 months–8 years) with lead intakes averaging 10.6 mg lead/kg bw/d (body weight/day). Absorption averaged 53% of intake, and retention averaged 18% of intake. Also, Ziegler et al. (1978) studied lead absorption by 12 infants (aged 14–746 days) whose lead intakes exceeded 5 mg/kg bw/d. These fractional absorption estimates, having been derived from studies in the 1970s (when exposures to lead were many times higher than current levels), may not be directly applicable to current estimates of kinetics. Still, until more data at lower exposure levels become available, it is appropriate to use distinctly higher fractional absorption estimates for infants. It is unclear over what age period in childhood do the lead absorption characteristics become more like those of adults than infants. Although studies specifically evaluating agerelated changes in fractional absorption of lead by children older than infants are not yet available, some insight might be drawn from other studies. Based on analyses of stable lead isotope profiles of nine immigrant children from Eastern Europe living in Australia, Gulson et al. (1997) found the fractional absorption of ingested lead by the children (aged 6–11years) to be comparable to absorption patterns of adult females in the 29–37 years age range. Whether the 40–50% absorption values for ingested lead obtained for such subjects typically under 2 years old apply to children in the 2–6 years age range remains unclear. Lower absorption values for 2–6 year old children are supported by data from Angle et al. (1995), who suggested that absorption of ingested lead among 2–3 year old children was 10–15%. See Mushak (1991) for more detailed discussion of factors (including both physiological and dietary) potentially underlying age-related differences in GI absorption of inorganic lead. 20.4.3

Influence of Nutritional Status and Dietary Factors on Lead Absorption

The influence of nutritional status and dietary factors on blood and on tissue lead distributions have been most clearly observed at low levels of lead exposure, for example, those more likely to exist currently in the post-2000 than the pre-1980 period. Since the mid-1980s, lead exposure has declined markedly in countries that have discontinued the use of lead solder in food and beverage cans and have phased out the use of lead-based gasoline additives. The beneficial effects of optimal nutrition are enhanced under these lower exposure circumstances (Mahaffey, 1995; Bogden et al., 1997), but the primary approach to effectively reduce lead impacts on public health is to limit human exposures. 20.4.3.1 Total Food Intake Adults in the fasting state have been reported to absorb a substantially greater fraction of lead compared with the fraction absorbed in the nonfasting state (Blake and Mann, 1983). However, there appears to be a lack of data regarding

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comparable information on the effects of fasting state on lead absorption by children or young nonhuman primates. 20.4.3.2 Calcium Dietary calcium can influence gastrointestinal absorption of lead through both acute and long-term effects of low dietary calcium intake. Numerous studies of experimental animals fed low calcium diets have established that calcium deficiency increased both tissue retention and toxicity of lead (Mahaffey-Six and Goyer, 1970). Longterm calcium deficiency produces a number of physiological adaptations. These include increased concentrations of various binding proteins and the stimulation of endocrine systems and regulatory systems, for example, 1,25-dihydroxycholecalciferol and parathyroid hormone. These secondary changes produced by calcium deficiency also affect the biokinetics of lead. Overall, calcium deficiency generally increases lead uptake. Experimental animal studies show that simultaneous ingestion of lead with reduced calcium concentrations in the incubation medium (i.e., comparable to a low calcium meal) enhanced absorption of lead. For example, Barton et al. (1978a), using ligated intestinal loop techniques to measure lead absorption, found that when the concentration of calcium in the incubation medium varied within physiological ranges, lead absorption decreased with increasing calcium concentration. Prior conditioning by low or high calcium diets did not significantly alter the rat’s lead absorption in vivo. Lower lead absorption was observed in rats and chicks (Smith et al., 1980) during studies on the role of vitamin D in lead absorption. Reduced lead uptake from ligated gut loops upon addition of calcium to incubation media was reported by Barltrop and Khoo (1975), and Meredith et al. (1977) found that oral calcium given immediately before lead very effectively decreased lead absorption in rats. Analogously, higher dietary calcium intake decreased lead absorption in humans (Ziegler et al., 1978; Blake and Mann, 1983; Heard and Chamberlain, 1982). Mechanisms that produce changes in lead absorption due to calcium status have become increasingly better understood (Fullmer, 1997). Aungst and Fung (1985) found that the apparent systemic availability of 1, 10, and 100 mg/kg oral lead doses were three- to fourfold greater in calcium-deficient than in control animals. However, the intestinal absorption of 10 kg/mg doses of oral lead was unaffected by calcium supplements. Such differences were thought to reflect the roles in lead and calcium absorption of dietary vitamin D (cholecalciferol) and metabolically active vitamin D (1,25-dihydroxycholecalciferol). Mykkanen et al. (1982) reported that in chicks, both cholecalciferol and 1,25-dihydroxy vitamin D3 affected both the 203 Pb and 47 Ca absorptive processes, but the nature of these responses were not identical, suggesting differences in the transport path or the macromolecular interactions of these metal ions (or both) during the course of absorption. Studies of lead conclusively verified specific, high-affinity binding of lead to several calcium-biding proteins and suggested that it may be a general property of certain intestinal calcium-binding proteins (Fullmer et al., 1985). Fullmer (1997) further investigated time courses and dose–response relationships for these interactions. By feeding five different levels of calcium and five levels of lead to chicks, Fullmer found that lead ingestion and calcium deficiency, either alone or in combination, generally increased serum 1,25-dihydroxy vitamin D levels over most of the range of dietary lead and calcium intake. However, with severe calcium deficiency, consumption of lead produced marked decreases in 1,25-dihydroxy vitamin D levels. Overall similarities in the responses of 1,25-dihydroxy vitamin D, intestinal calcium absorption, and calbindin-D indicate that the predominant interaction between lead and calcium is mediated via changes

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in circulating 1,25-dihydroxy vitamin D concentrations, rather than directly through the intestine. Kidney and bone lead levels also changed in response to these dietary manipulations, suggesting that added effects occur that do not fully depend on the concentrations of 1,25-dihydroxy vitamin D, although this appears to be the predominant control mechanism for intestinal absorption. 20.4.3.3 Iron Iron deficiency increases lead tissue deposition and toxicity (Mahaffey-Six and Goyer, 1972). Ragan (1983) demonstrated sixfold increases in tissue lead in rats when body iron stores were reduced but before frank iron deficiency developed. Also, Hamilton (1978) and Flannagan et al. (1979) reported significantly increased absorption of lead from the GI tract of iron-deficient animals. Lastly, based on results obtained by in situ ligated gut loop techniques, Barton et al. (1978b) reported that iron deficiency (secondary to bleeding and to iron-deficient diets) increased lead absorption and that iron loading decreased lead absorption. Ferritin has been shown to bind lead both in vivo and in vitro. In rats fed an iron-deficient diet, the ferritin concentration was low, permitting increased transfer of lead to blood rather than retention in the small intestine bound to ferritin. Transferrin is increased in irondeficiency anemia as a result of increased synthesis. Although transferrin binds iron preferentially, transferrin also transports a number of trivalent and divalent cations such as plutonium, americanium, chromium, cobalt, manganese, and copper, among others. A protein that specifically binds lead, as well as iron, was isolated from both the rat and from the human duodenal mucosa (Conrad et al., 1992). The influence of iron status on lead absorption has also been studied in human subjects, with mixed results (Flannagan et al., 1979; Watson et al., 1980). To date, it is not clear whether these mixed results reflect differences in the severity of iron deficiency, differences in analytical approaches, or other undefined factors. Despite the lack of clarity regarding mechanisms, iron therapy is proving to be a valuable adjunct in the treatment of low-level lead toxicity (Granado et al., 1994). 20.4.3.4 Influence of Chemical Forms of Lead on Gastrointestinal Absorption The effect of the chemical forms of lead on gastrointestinal absorption of the metal can be described only in general terms. Lead bound to alkyl compounds is readily absorbed and concentrates in tissues high in lipid, such as the brain. The percentage of absorption and tissue distribution of alkyl lead differs markedly from those of inorganic lead compounds. Among inorganic lead compounds, the particle size of ingested lead plays a major role in determining the fractional absorption. Barltrop and Meek (1979) showed a fivefold enhancement in lead absorption by rats when the particle size of lead was reduced from 196 to 6 mm. Healy et al. (1982) reported that lead sulfide (considered to be one of the least soluble lead compounds) had increased solubility in gastric fluid (apparently as a result of chemical conversion to the more soluble chloride) when the particle size was reduced from 100 to 30 mm (Healy et al., 1982). Information about the influence of the chemical form of lead on its absorption is frequently complicated by limited information on the experimental conditions, including other factors in the diet, particle size of the inorganic lead source, physiological condition of the animal, and so on. In vitro solubility does not appear to predict well the degree of in vivo absorption (Sartorelli et al., 1985). Despite this limitation, the use of multiple in vitro methods to estimate solubility continues. Interpretation of results from these in vitro methods is further complicated by organ-to-organ differences in bioaccumulation of lead. For example, when immature swine were fed two fully characterized soil samples from a western U.S. Superfund

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site (Casteel et al., 1997), the bioavailability ranged from about 50–90%, depending on the organ system used to express dose (e.g., blood, liver, or renal lead concentrations). In mechanistic terms, lead absorption depends on such factors as chelation, membrane permeability, solubility, and particle size (Brezinski, 1976; Huisingh and Huisingh, 1974). The coordination chemistry of Pb2þ is likely to play an important role in many of these factors. Coordination with proteins may be a determining factor for the availability of lead for absorption and transfer across the mucosal cell. It is important to know in which form(s) lead is (are) available and what ligands exist in mucosal cells that may be vehicles for absorption or inhibition. Metals can precipitate or coagulate proteins in solution. Also, metal salts often show increased solubility in body fluids (Fairhall, 1924), lead carbonate being, for example, about 300 times more soluble in serum than in water. Such metal–protein interactions depend on factors such as the radius, charge, and coordination number of the metal, as well as factors intrinsic to the proteins, such as size and basicity. In a protein-rich environment, the local metal ion power and complexing ability of the food should be considered. 20.4.4

Absorption Following Inhalation

Absorption of lead from the pulmonary system depends, in a major way, on the particle size of the inhaled lead. It is difficult to determine what fraction of lead dust in inhaled air actually gets deposited in the gas exchange airways and taken up by alveolar cells or what fraction is deposited on the conducting airways and is eventually passed out through the trachea and swallowed with mucus from the trachea or nasal passages. It is clear, nevertheless, that the pattern of lead deposition in the respiratory tract is affected by the particle size of the inhaled aerosol and the ventilation rate (Chamberlain, 1983). The rate of absorption of lead from the particles deposited also depends on solubility of the chemical species of lead. In humans, the absorption of lead from the lung is usually rapid and complete within 24 h. Apparently, even relatively insoluble lead compounds can be taken up directly into the general circulation in this way. Ligand exchange ability is also a key factor here, since it relates to physiochemical properties of the available lead species and the surface properties of the particles involved. All species of lead compounds deposited in the deep (alveolar) lung region are thought to be more or less completely absorbed into the blood stream (Morrow et al., 1980), with distinctly greater alveolar deposition typically occurring for particles less than 2.5 mm in diameter than for larger-sized particles. However, as is often seen with various environmental or occupational exposures, it is not unusual for lead dusts in inhaled air to contain lead particles of large enough size (larger than 2.5 mm diameter) so that many are cleared from the conductive airways by mucociliary action. Lead particles larger than 2.5 mm in diameter deposit mainly in the nasopharyngeal and tracheobronchial regions (which constitute the upper repiratory tract) where they can be transferred by mucociliary action and then swallowed. Therefore, because many such particles are swallowed, factors that affect the GI absorption of lead can also often significantly influence the bioavailability of inhaled lead. Thus, depending on the particle size and concentration of lead associated with a particular air exposure source, the digestive tract can also be an important avenue of lead absorption following inhalation. Chemical speciation of lead dust in occupational settings has shown marked variability in the size of particles generated in primary smelters (Spear et al., 1998). Depending on the process performed (e.g., samples from ore storage, sintering, or blast and dross furnaces), the particle size, mineralogy, and extractability of the constituent lead can differ substantially, and, consequently, changes to the ore during processing are thought to influence the biological availability of the lead.

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771

Dermal Absorption of Lead

Skin absorption is not usually considered to be a significant mode of lead uptake (Minot, 1929), unless the metal is present in its more lipid-soluble organic forms (such as tetraethyl lead). Florence et al. (1988) found that inorganic lead can be absorbed through skin and rapidly distributed throughout the body. Of note was the observation that the distribution tends to vary in some ways from that of ingested lead. For example, skin absorption gave rise to increased lead excretion in sweat, although similar increases in blood and urine were not observed.

20.5 DISTRIBUTION Lead is absorbed into blood plasma, where it rapidly equilibrates with extracellular fluid. More slowly, but within minutes, lead is transferred from plasma into blood cells (Chamberlain, 1985; Simons, 1986). The typical concentration of lead in whole blood is about 10 6 M. Because 95–98% of the lead is bound in red blood cells, about 10 8 M is present in the plasma. If the distribution of lead between plasma and cytosol is similar to that of calcium (10,000:1), the cytosolic concentration of lead in exposed individuals should be in the picomolar range. In animal experiments, no constant relationship has been found between the lead concentrations in blood and in soft tissue. Thus, as earlier noted by Kazantzis (1988), controversy exists with regard to whether lead in blood represents biologically active lead, and indeed, the extent to which the two may be linearly related. Improved detection limits for analytical methods have increased our ability to determine concentrations of lead present in plasma. Concerns remain, however, that even very slight hemolysis of erythrocytes during the separation process can transfer lead into the plasma fraction. Consequently, data on plasma lead concentrations must be treated with caution unless the technique can establish that what is thought to be plasma lead does not simply represent the in vitro transfer of lead from erythrocytes. From the blood plasma, absorbed lead is distributed to different organs, with liver and kidney attaining highest concentrations. Of interest, the peripheral nervous system may accumulate much more lead than the central nervous system (CNS). Also, marked variation occurs in lead distribution within various other tissues and organs (Barry, 1975, 1981; Drasch et al., 1987; Drasch, 1974, 1997; Drasch and Ott, 1988). In this regard, there are several important corollary observations. First, lead tends to accumulate wherever high calcium levels are found. Highest lead concentrations are, therefore, found in bone, especially in dense cortical bone. Second, within and among soft tissues, highest lead concentrations seem to accumulate in those tissues and organs having the highest mitochondrial activity. Likewise, within a given organ, the highest concentrations occur in regions with the highest mitochondrial activity (e.g., in the renal tubule and in the choroid plexus and cerebellum of the central nervous system). The skeleton contains more than 90% of the body burden of lead when measured at steady state. However, this pool is neither homogeneous (Kehoe, 1961; Chamberlain, 1985; Rabinowitz et al., 1976) nor is it static (Gulson et al., 1995). This association with bone is related to lead’s similarity to calcium and formation of insoluble lead phosphate. As described subsequently, there are many bone pools of lead. Bone is a very complex organ, having varying density and structure, depending on skeletal site and function. The turnover of tissue lead is high throughout life. Among young adult women, between 50 and 75% of

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blood lead reportedly comes from tissue stores (i.e., skeletal) rather than the current environment (Gulson et al., 1995, 1996). Under conditions of physiological stress for calcium (including pregnancy and lactation), the release of bone lead becomes even higher (Gulson et al., (1998a, 1998c); that is, the initial source of the lead is environmental lead that has been accumulated in tissues over previous years. Estimates of the fraction of current blood lead derived from bone were also provided by Smith et al. (1996), who reported that the skeleton contributed 40–70% of lead in blood among five subjects who had trabecular bone samples obtained at surgery. There are now thought to be three basic types of bone lead pools:(1) rapidly exchanged lead in very metabolically active portions of bone; (2) lead in trabecular or spongy bone; and (3) lead in dense cortical bone. Lead turnover in these three basic types of pools appears to roughly parallel the relative rates of calcium turnover. Various observations suggest marked variation in the distribution and turnover rates within these pools, at least partially depending on the particular region of the skeleton studied. It is important to note further that although lead concentration is lower in trabecular than in cortical bone, the mass of trabecular bone is, on average, four times that of cortical bone and it is approximately four times as labile. Therefore, trabecular bone may represent a much more metabolically important pool of lead. The distribution of lead in tissues reflects a state of constant, dynamic equilibrium. As noted in the section on excretion, many methods of enhancing lead excretion are also influenced by lead’s redistribution within the body (Cory-Slechta et al., 1987). Clearly, any situation that mobilizes the very large, relatively stable pools of lead within the body, particularly those in the bone, will lead to the redistribution of lead to a variety of tissues. This redistribution is thought to explain the increased symptomatology that is frequently noted in lead-poisoned children during acute illnesses. Redistribution is known to occur during pregnancy and, even under usual circumstances, can result in increased risk to the fetus, particularly in women with prior lead poisoning. There is also some evidence that the osteoporosis of aging may be accompanied by the significant mobilization of lead from bone pools. It is clear that much additional information is needed to clarify more fully the various physiological and pathological conditions of enhanced mobilization and redistribution of lead. 20.5.1

Excretion

Lead is excreted from the body mainly by urinary and fecal routes, with fecal excretion representing the sum of unabsorbed endogenous lead from saliva, bile, and (to a lesser extent) other gastrointestinal secretions, plus the unabsorbed portion of inhaled and ingested lead. Less excretion occurs through sweat and integumentary losses (including skin, hair, and nails), and these routes account for only a small portion of total excretion. Under conditions of fairly constant exposure to low lead concentrations, a steady-state condition evolves, wherein excretion approximates intake (Rabinowitz et al., 1976) and 70% of intake is excreted via urine. Under short-term conditions of low-level increased exposure (Chamberlain, 1985), 60% is retained by the body and 40% excreted. In a 14-day study of human volunteers receiving a single dose of 203 Pb, 18% of the dose was excreted whereas 35% was retained in blood, with a residual of 47% in soft tissues and bone storage. Extrapolating this result would predict an eventual 30% excretion and 70% retention of this single dose. Urinary excretion of lead is quite complex, depending on the situation, but is most likely a function of plasma lead levels. Under most conditions of exposure to relatively low lead

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concentrations (e.g., at blood lead levels below 25 mg/dL), the concentration of lead in plasma is very low (about 0.01 mg/dL) and not related to whole blood lead. Above this concentration, plasma lead increases significantly, as does urinary lead excretion. At blood lead concentrations below 25 mg/dL, urinary clearance of lead has been estimated at 1.1% per day. As blood lead rises above this, urinary clearance appears to rise at a rate reasonably related to the increase in plasma lead (Chamberlain, 1985). At very high blood lead concentrations, renal dysfunction may decrease urinary excretion of lead. In considering fecal lead excretion, one must differentiate between fecal lead that is unabsorbed from ingestion or inhalation and fecal lead that truly represents endogenous fecal excretion. Endogenous excretion was measured by Rabinowitz et al. (1976) using stable isotope studies and by Chamberlain (1985) after inhalation and/or parenteral administration of 203 Pb. Endogenous excretion is often estimated by comparing renal clearance to apparent total body clearance. All such estimates suggest a clearance of 0.5% per day at blood lead levels 525 mg/dL. It is likely that this rate of clearance is basically independent of blood lead and so it does not significantly increase with increasing blood or plasma lead. Excretion by all other routes is at a rate of 0.2% per day and is, again, essentially independent of blood lead concentration. The total for all these excretion routes is 1.8% per day at blood lead concentrations less than 25 mg/dL and somewhat greater at higher blood levels because of increased urine lead excretion at higher concentrations. A special form of excretion is that which occurs via breast milk. Various studies of maternal breast milk composition indicate that breast milk lead concentrations appear to correlate well with maternal blood lead concentrations. One report of plasma lead concentrations in mice (Keller and Doherty, 1980a, 1980b) suggests that breast milk lead concentrations are more closely related to plasma lead concentrations and can be as much as 25 times that of plasma lead. This suggests that at lower blood lead concentration in the mother, breast milk would likely represent a minor route of excretion and is usually a minor exposure route for the infant. However, at higher blood lead concentrations with increasing plasma lead concentrations, it is plausible that a significant amount of lead could be mobilized from the maternal skeleton in lactating women and that breast milk could represent an important exposure pathway for breast-fed infants. In fact, using stable isotope methods, Gulson et al. (1998c) demonstrated this mobilization among women with blood lead concentrations less than 10 mg/dL. Breast milk lead appears to be linearly related to blood lead and has concentrations similar to plasma. For blood lead levels in the range of 2– 34 mg/dL, breast milk contains les than 3% of the quantities of lead in blood. The amounts of lead released from the skeleton show much person-to-person variation, suggesting that among women who have had substantial prior lead exposure, it is important to assess this as a possible exposure source for the infant. A variety of specific methods have been used to alter the clearance of lead. Aub et al. (1925) showed that urinary excretion could be enhanced by acidification, presumably by mobilization of lead from relatively stable pools. After some initial enthusiasm for the therapeutic potential of this intervention, it was abandoned because such therapy often also enhanced symptomatology, presumably by redistributing lead to soft tissues. A variety of chelating agents have also been used to enhance the clearance of lead. The majority of chelating agents, including ethylene diamine tetraacetic acid (EDTA), dimercaptosuccinic acid, and d-penicillamine, enhance clearance by binding lead and promoting urinary clearance. Dimercaprol is another reasonably effective chelator of lead, but it predominantly enhances biliary excretion of lead.

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20.6 KINETICS Basic to an understanding of the effects of lead exposure on animals and humans is an appreciation of the kinetics of lead in living animals. This requires recognition and knowledge of the various phases of lead kinetics, including absorption, distribution, and clearance. Although lead may exist in the body in various ionic forms and compounds, it is not properly considered, as a material, to be “metabolized” by the body. Rather, it is transported by various more or less metabolically active complexes and compounds. Useful terms to describe this distribution are lead kinetics and biokinetics. Several diverse approaches have been used to assess the lead kinetics in mammals. Of interest is the fact that all of the approaches have produced similar conclusions. The most important of these is that delineation of lead kinetics requires a multicompartment model, with some compartments being rather large and relatively stable, whereas other compartments are smaller and comparatively labile. Also, it appears most likely that the kinetics of lead at lower concentrations of exposure may be considered linear, whereas at high exposure concentrations, they appear to be nonlinear. This has potentially important implications in considering the biological effects of lead at varying exposure levels. Several types of approaches have been used to investigate the biokinetics of lead. Such approaches mainly include the following: 1. Exposure of both experimental animals and human subjects to lead orally or via inhalation, wherein increasing amounts of lead are introduced and the resulting accumulation and excretion of lead are measured by a variety of methods. 2. Introduction of radioactive or stable isotopes of lead to determine kinetics without increasing exposure and disturbance of the steady-state situation. 3. Study of the spontaneous clearance of lead in situations where exposure to a high concentration of lead has been terminated. 4. Study of the distribution of lead in various tissues by postmortem examination. The last approach has been very useful in animal studies but little used in studying lead distribution in humans. It is important to remember that there is notable interspecies variation with regard to the kinetics of lead due to many factors, for example, diet, physiology, and relative tissue mass. For these reasons, it is clear that studies of animals other than humans must be viewed with great caution in attempting to understand human kinetics. Since the greatest concern is with human exposure, human studies are emphasized below. The earliest studies of lead balance in humans and animals were by Aub et al. (1925), who used several of the above-noted approaches in a series of studies of exposures of animals to lead by both inhalation and ingestion. Aub concluded from those studies that lead was most readily absorbed from the respiratory tract, particularly by the inhalation of “finely divided particles,” and he noted that lead was somewhat less well absorbed by ingestion. In his animal studies, Aub also measured lead concentrations in a variety of tissues after exposure to determine the distribution of lead in the body. In addition, he performed autopsies on human bodies to evaluate the internal distribution of lead in lead-exposed patients. Aub concluded that especially after termination of lead exposure, all the lead is “permanently” stored in bone and, accordingly, that lead was harmless to a person unless there was “recent absorption from an external source or mobilization of a skeletal store.” He conducted

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extensive further studies of lead excretion in animal and human subjects and found spontaneous excretion to be very low and variable, but noted that calcium-deficient diets and administration of acids markedly increased lead excretion (especially if both factors are applicable at the same time). Kehoe (1961) subsequently conducted a landmark series of experiments that involved long-term exposure of human volunteers to lead by inhalation and ingestion. The subjects were observed prior to and subsequent to exposure to lead to determine both baseline lead balance and balance during a “recovery” stage from lead exposure. Kehoe’s studies were summarized by Gross et al. (1975), who made several important observations that are relatively consistent with Aub’s earlier findings. For example, there was considerable variability of observations within the subjects during the control period, which appeared to relate, at least in part, to variability in dietary lead exposure. Also, these variations in ingested lead (about a mean of 191 mg per day) were paralleled by variations in fecal and urine lead with an overall net negative lead balance (in most subjects), as calculated by intake versus urine and stool output. It is not clear whether this reflected an actual net negative balance or was simply a result of not being able to measure airborne lead exposures during the control period. A later study by Rabinowitz et al. (1976) substantially clarified our understanding of absorption and distribution of lead in the body. The results of this study, which are among the most carefully obtained data available, have been extensively used by the subsequent kineticists. Rabinowitz et al. (1976) studied “normal” volunteers under standard conditions of diet and activity. These volunteers were fed a low lead diet supplemented to approximate their usual level of lead intake by addition of 204 Pb as a tracer. Since this isotope is rare in most usual sources of exposure, it provided a stable tracer for purposes of defining the kinetics of lead in the body. Lead isotope distributions in samples were determined by mass spectroscopy. These adult volunteers had prestudy blood lead concentrations of 16–25 mg/dL and had intakes of lead between 156 and 215 mg per day during the study, which approximated their prestudy intake. One of the five subjects was studied for 10 days. The other four were studied for longer periods, ranging from 108 to 210 days. The range of absorption observed was 6.5–13.7% of the ingested dose. Of this absorbed lead, 54–78% was excreted in the urine, with most of the rest being excreted via bile and integumentary losses. The translocation of the tracer lead was best described by a three-compartment kinetic model. The first compartment, which included 1.5–2.2 times the amount of lead in blood, contained an average of 1900 mg of lead and was turned over in about 36 days. A second one, which comprised most of the soft tissue lead, contained about 600 mg of lead and was turned over every 40 days. These authors noted, however, that the total of less than 3 mg of lead in these two labile pools was much less than the 10–30 mg found in autopsy studies, suggesting that most of the soft tissue lead must have been in a more stable compartment. The third, large and comparatively stable, compartment was primarily comprised of bone lead. In these subjects, it contained about 200 mg of lead and was turned over approximately every 104 days. Of interest with regard to this pool was the comparison of total lead and tracer lead to total lead ratios in cortical and trabecular bone from the iliac crest. The total lead concentration in the cortical bone by weight was approximately twice that in trabecular bone. However, trabecular bone had a two- to threefold greater ratio of tracer to total lead, suggesting that it was turned over much more rapidly than the cortical bone. Of further interest was their calculation, based on these measurements, that the iliac bone received lead three–seven

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times more rapidly than did the very stable pool as a whole. These observations make it clear that bone cannot be regarded physiologically as a single pool. Furthermore, Rabinowitz et al. (1976) emphasized that all of the pools are in dynamic equilibrium with each other and, therefore, any changes in the movement of lead from one to another can cause significant changes in the measured amount. Factors that might cause movement of lead from the large stable pool into blood, and hence to target soft tissues, are of particular interest. Steenhout (1982) developed a kinetic model of lead distribution based on data regarding lead in teeth in children and adults in three regions of Belgium. Her model, basically consistent with the Rabinowitz model, suggests that the rate of transfer of lead to teeth is 1.85 ppm/yr/mg per 100 mL blood. According to the model, lead accumulation in teeth and dense (cortical) bone is linear and continuous over age, suggesting very slow loss from dense bone (approximately 0–0.005 ppm/yr/mg per 100 mL blood). In contrast, estimates of lead loss from “porotic” bones, such as ribs and vertebrae, are on the order of 0.06 ppm/yr/mg per 100 mL blood. Steenhout concluded that the apparent nonlinearity of lead transfer in some other studies reflected this relatively rapid loss of lead from such porotic bone. She also suggested that her data support the concept that for dense bone, there is no loss of lead with increasing age and, therefore, the osteoporosis of age should not represent a risk for lead mobilization. Chamberlain (1985) used data from a variety of data sets, including some from other investigators and some he had developed to assess several aspects of lead kinetics in humans. He focused primarily on volunteer feeding experiments to observe the response of blood lead to either airborne exposure or to dietary intake. His discussion concerned largely inorganic lead and relatively short-term studies. He found that lead is absorbed rapidly into plasma and then into extracellular fluid in minutes, based on experiments with the injection of radioactive lead tracers. He noted that in a time frame measured in tens of minutes, lead in plasma, and lead from extracellular fluid via plasma, becomes largely bound to red blood cells. Approximately 58% of a dose of lead was found to be bound to red cells after 20 h. Chamberlain (1985) also found that excretion of lead after a single dose occurs over a month and that lead storage in tissues and bone persists for months to years. He further noted that the accumulation and distribution of lead differ in several ways from those of strontium. Most importantly, the attachment of lead to red cells appears to retard, rather than to promote, the distribution of lead to storage sites. Chamberlain’s autopsy studies showed that relative to “dose,” there is more lead than strontium or calcium in soft tissues. All the studies Chamberlain reviewed agreed that the transfer of lead to excreta from blood occurs over a period of about one month. He also noted that at low-level exposures, urinary excretion is two or more times greater than excretion via stool. In his discussion of transfer of lead to bone, he analyzed both the discrepancies and consistencies among available data sets. Of most importance for kinetic modeling is the observation in some of the data sets that the lead concentration in trabecular bone is similar to that in cortical bone, whereas in other data sets the lead concentration in trabecular bone was much higher than in the cortical bone. Some of this discrepancy may be related to the duration of the studies, since short-term (versus longer) studies may show relatively more lead in the presumably more labile trabecular pool. Chamberlain’s resorption rates from storage in bone are inferred indirectly, that is, based on studies of strontium turnover, given the assumption that the rates do not significantly differ for various trace minerals. However, considering his review of differences between blood, plasma, and tissue distribution of strontium and lead, this assumption may not be entirely valid.

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Based on these studies, Chamberlain (1985) suggested a mean life of lead to be 12.5 years in trabecular bone and 50 years in cortical bone. His estimate for the mean life of lead in soft tissues, derived from autopsy data, was 500 days. The relationship between urinary clearance and blood lead appeared to be constant in the range of “normal” blood lead concentrations, but to increase proportionately at blood lead levels above 20–25 mg/dL. This implies a decreasing apparent relative uptake with increasing blood lead concentration. Chamberlain reviewed some data on intestinal absorption, noting that uptake of soluble lead tracers is markedly affected by a period of fasting, with an average (in several studies) of 8% uptake when lead was taken with a meal and 60% when taken after an overnight fast, if the fast is continued for several hours after the lead ingestion. He noted that insoluble lead sulfide absorption was less affected by fasting (12% absorbed in fasting versus 6% with meals). He also reviewed data showing that the addition of calcium and phosphorus salts markedly decreases the absorption of soluble lead. Chemical incorporation of lead with foodstuffs did not alter lead absorption below the levels observed when lead was administered with another metal. The discussion of airborne lead exposure by Chamberlain (1985) focused on three major factors of exposure to inorganic lead only: (1) airborne lead concentration; (2) ventilation rate/volume; and (3) fractional distribution of the aerosols. He did not consider airborne exposure to organic lead. The particle size in the inhaled aerosol and the “residence time” in the pulmonary region, determined largely by respiratory rate, appeared to be relatively unimportant. Once retained in the lungs, the deposited lead is essentially completely absorbed into the bloodstream within 24 h. Of note is that many larger particles, deposited higher in the conductive airways, are returned by mucociliary clearance to the pharynx and thereafter ingested, reducing this mainly to the case of ingested inorganic lead. A somewhat more complex model of lead distribution was suggested by Bernard (1977). His “reference man” had a total body burden of lead of 120 mg. Of this total burden, 110 mg was in bone and the rest in soft tissues. His model proposed at least two bone pools, a slow pool in cortical bone and a relatively labile pool in trabecular bone. Bernard further proposed two soft tissue pools, one of which is relatively large and slow and another is quite large and very rapidly turned over. Although this model is logical and based on experimental observations, its actual validity is somewhat subject to question because it is based exclusively on studies in rats and nonhuman primates, which may markedly differ physiologically from humans. Subsequently, Schutz et al. (1987a) observed the decline of blood lead after termination of occupational exposure by studying two separate and somewhat different groups. The first group included workers who no longer worked in the lead industry, whereas the second included those who were removed from work due to a blood lead increase to above 3 mmol/L (60 mg/dL). The first group was older, had longer periods of exposure, and generally had lower mean blood lead levels than the second. The subjects also had bone lead concentrations estimated by the use of X-ray fluorescence (XRF) of the middle phalanx of the left index finger. A two-compartment model was found by Schutz et al. (1987a) to provide a satisfactory fit to his data, with the “fast” compartment having a half-life of 30 days and the “slow” compartment a half-time of 5.6 years. There was notable intersubject variation, which was suggested to represent “considerable variation in risk at a given exposure level.” The “slow” pool was turned over somewhat more rapidly than other reported rates. The bone lead observations by X-ray fluorescence correlated positively with estimates of the slow pool, but the coefficient of correlation was rather low (r ¼ 0.36). Schutz hypothesized that the slow pool may actually consist of a combination of two bone pools, one of trabecular and

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the other of cortical bone lead. This hypothesis was used to explain why the measurement of bone lead somewhat differed between the two groups. He suggested that this difference may relate to differing proportions of lead in the cortical versus trabecular bone, the concept being that long-term exposure would result in higher relative levels in cortical bone than in trabecular bone. If this is correct, it has implications for considering widespread application of noninvasive methods of bone lead measurement to research and clinical assessments. The biological redistribution of lead pools within the body has been described in a series of reports based on the evaluation of differences between stable isotope profiles of body lead compartments accumulated in separate geographic locations and during different life periods. Two persons had earlier been investigated by Manton (1977, 1985 in initially describing the process. Further research to confirm these changes, as well as to verify their occurrence in a number of subjects, was conducted by Gulson et al. (1995, 1997, who investigated a cohort of adult women who immigrated to Australia from Eastern Europe and Russia during the early 1990s. These subjects had accumulated tissue stores of lead in Europe that had a stable isotope ratio distinctly different from that of lead in Australia. Such differences enabled the researchers, by means of meticulous measurement of stable isotope ratios via thermal ionization mass spectroscopy, to identify the proportion of blood lead from the contemporaneous environment and that mobilized from tissue lead stores accumulated earlier in Europe. The data obtained showed that among these young adult women, 45–70% of lead in blood came from the long-term tissue stores, presumably in bone (Gulson et al., 1995). These proportions occurred at blood lead concentrations that averaged 55 mg/dL as the result of environmental lead exposures typical of developed countries where steps had been taken to restrict lead exposures. The above study also aimed to determine the influence of pregnancy and lactation on this mobilization. During pregnancy, blood lead concentrations in these female subjects increased by about 20%, on average, with individual changes ranging from 14 to þ83% (Gulson et al., 1997). Among those subjects whose blood lead levels increased during pregnancy, the mean increase in mobilization of long-term tissue lead stores varied from 26 to 99%, averaging about 30% (Gulson et al., 1997). Skeletal lead mobilization continued to be elevated after the pregnancy. Observations for infants born to these mothers showed that the long-term tissue stores of lead in the mothers had been transferred to the fetus and that among those infants who were breast-fed, additional transfer of lead continued to occur during breast-feeding (Gulson et al., 1998a, 1998c). The transfer of maternal skeletal lead to the fetus as shown by stable isotope analyses has also been confirmed among nonhuman primates (Franklin et al., 1997). Barry (1985) measured lead concentrations in tissues of 129 subjects at autopsy and presented extensive data on lead content in various tissues from this very large series. He noted that consistent with other studies, the content of lead in bones was much higher than in soft tissues and that the levels of lead in dense bone were much higher than in more cancellous bone. For example, petrous bone had the highest levels and ribs the lowest (by an approximate ratio of 4:1), a finding of much importance in understanding the generally nonhomogeneous distribution of lead in bones. Barry noted that in those soft tissues with the higher lead amounts, the concentrations in males exceeded those in females by about 30%. He stated that soft tissue lead concentrations increased with age only through the second decade of life and were thereafter stable. Children were reported to have soft tissue concentrations similar to adult females, but had much lower bone concentrations. He also stated that in adults, over 90% of the lead was in bone, with more than 70% being in dense bone. Among occupationally exposed adult males, 97% of their total body lead was in bone.

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Barry noted that the increase of lead in bone with age, with respect to stable soft tissue levels, is consistent with the hypothesis that lead in bone is not available to soft tissues. He also indicated that the lack of decline in the bone lead concentrations in the face of demineralization with increasing age suggested that lead was not mobilizable even under conditions of massive calcium turnover. However, it must be further noted that Barry expressed bone lead changes with age only as concentration and not as total lead, and he ignored the fact that demineralized bone would have decreased total mass. Thus, the apparent constant concentration of lead in bone may actually reflect a markedly decreased total amount of lead in bone. Predicting quantities of lead mobilized from bone requires bone lead concentration data for both cortical and trabecular bone. Data sets providing bone lead concentrations among adults and children described above were obtained during time periods in which environmental lead exposures were much higher than now present in many developed countries. Such lower lead exposures result in distinctly lower bone lead levels. For example, Drasch and coworkers obtained data on bone lead concentrations for cases coming to autopsy in Munich between the early 1970s and 1994 (Drasch, 1974, 1997; Drasch et al., 1987; Drasch and Ott, 1988). These data are for subjects living in the same geographic vicinity in southern Germany. Between 1974 and 1994, trabecular bone lead decreased from 2.5 mg/kg (1974) to 1.7 mg/kg (1984) to 0.7 mg/kg (1994). Compact bone decreased from 5.5 mg/kg (1984) to 2.8 mg/kg (1994). These data are for adults. Changes in bone lead can be expected to be even more dramatic among young children, who (unlike adults) do not have long-term stores of lead accumulated during decades of much higher lead exposures. Drawing upon the above types of advances in regard to lead kinetics and associated modeling thereof, substantial further progress has been made during the last several decades in developing and refining biokinetic model systems to project likely increased risk of lead-related toxicity in human population groups due to various lead exposure scenarios. Such biokinetic models involve stipulation of mathematical relationships among biological processes (e.g., absorption, distribution, redistribution, clearance, elimination) that determine variations in internal concentrations of the metal and associated potential for causing toxic effects. In particular, during recent decades, much progress has been made in developing various biokinetic models for predicting “blood lead” as the most widely accepted internal biomarker (discussed subsequently) traditionally used to index lead exposure/dose and to gauge consequent potential for leadrelated pathophysiological responses to occur. Some salient examples of progress made since the early 1980s in such modeling efforts include those of Marcus (1985a, 1985b, 1985c), in which available data sets were used to derive multicompartment kinetic models for lead, and that of Bert et al. (1989) who developed a compartmental model for adult males. Also, Leggett (1993) published an age-specific biokinetic model for lead that was developed originally for the International Commission on Radiological Protection (ICRP) but was later expanded to include additional features of use in consideration of lead as a chemical toxicant. In it, the transport of lead between compartments was assumed to follow linear, first-order kinetics, provided that the concentrations of lead in red blood cells remained below a nonlinear threshold level but a nonlinear relationship between plasma lead and red blood cell lead was modeled for concentrations above that level. Several other physiologically based models for bone-seeking elements published by O’Flaherty (1993, 1995, 1998 utilize information about age dependence of bone formation rate and take into account increasing localization of bone remodeling activity with age.

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In addition to the above models, the U.S. Environmental Protection Agency developed a widely used “Integrated Exposure/Uptake/Biokinetic” (IEUBK) Model for Lead in children (U.S. EPA, 1994a, 1994b; White et al., 1998; Hogan et al., 1998). As noted in Chapter 4 of U. S. EPA (2006), the IEUBK Model simulates lead exposure and biokinetics from birth to age 7 years and predicts quasi-steady-state average blood lead concentrations corresponding to daily average lead exposures averaged over periods of one or more years. Comprised of four subcomponent models (an exposure model, an uptake model, a biokinetic model, and a blood lead probability model), the IEUBK Model (1) calculates average daily intakes of lead (mg/ day) for each inputted exposure concentration (or rates) of lead for different multimedia lead exposure routes (via air, water, diet, dust, soil); (2) next converts the media-specific lead intake rates (calculated from the exposure model) into media-specific time-averaged rates of uptake (mg/day) of lead into blood plasma as the central compartment; followed by (3) biokinetic modeling that simulates transfer of absorbed lead between blood and other body tissues (bone, brain, kidney, etc.), lead excretion from the body (via urine, feces, skin, hair, nails), and predicts an average blood lead concentration for the exposure time period of interest; and (4) lastly utilizes a blood lead probability submodel that applies a log-normal distribution (with specific geometric mean/geometric standard deviation parameters) to predict probabilities for the resulting occurrence of a specified blood lead concentration in a population of similarly exposed children. Further efforts are now underway by the U.S. EPA to develop to an “All-Ages Lead Model” (AALM) that (1) aims to simulate lifetime lead exposure and biokinetics of lead in humans from birth to age 90 years and (2) expects at some near-future time to include a pregnancy module that simulates transplacental transfer of lead from mother to fetus. The AALM is expected to be capable of predicting lifetime exposure impacts on (a) the internal distribution of lead to bone and various soft tissues (brain, kidney, etc.) in addition to (b) blood lead distributions for population groups from birth up to 90 years of age. Chapter 4 of U.S. EPA (2006) summarizes the most salient features of the IEUBK Model, the AALM, and other biokinetic models alluded to above, and it notes available evaluations of the relative accuracy of the models in terms of how well their predicted blood lead distributions (or means) approximate actual blood lead distributions (or means) observed for modeled lead exposure scenarios. Models such as those discussed above have had varied success in predicting blood lead concentrations. Typically, such models have been more successful in predicting mean/ median blood lead concentrations than the overall distribution of blood lead levels. Dose-dependent differences in fractional absorption and distribution of lead complicate application of these models. Person-to-person variabilities (intensity of hand-to-mouth activity, nutritional status, etc.) can also modify the relationships between external (environmental) doses of lead and internal (blood, bone, soft tissue) lead concentrations. Among those models that recognize the importance of bone lead contribution to blood lead, the very limited available data on contemporary bone lead concentrations remain an important factor that needs to be more fully addressed. Workshops have been held that have attempted to reconcile differences between the modeled distributions of blood lead data and the observations from epidemiological studies. The availability of pooled analyses of epidemiological data from childhood lead studies in the United States has identified lead-contaminated dust loadings within the residence as a very strong predictor of blood lead among children (Lanphear et al., 1998). Further, the child’s age, race, mouthing behaviors, and study- or site-specific factors are influential in predicting blood lead at a given level of lead exposure (Lanphear et al., 1998).

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Analysis of available information on the kinetics of lead reveals some problems in using the currently available models to study lead in the living mammalian subject. Many of the models fit observed data for lead absorption, distribution, and excretion reasonably well and often describe observed changes in blood lead quite well. However, currently available models are generally not fully adequate for purposes of anatomic/physiologic description and prediction, thus limiting their usefulness in devising experimental models or developing useful approaches for clinical diagnosis/management of lead intoxication. For example, the assertion that bone lead is a single homogeneous pool, or two relatively homogeneous pools comprised of cortical and trabecular bone, does not comport well with certain available data. The Rabinowitz models and others derived from his data suggest that the most stable pool, thought to be largely bone, had a mean life of 30–50 years (Chamberlain, 1985; Simons, 1989; Kazantzis, 1988; Kehoe, 1961; Rabinowitz et al., 1976; Cory-Slechta et al., 1987). However, when studied directly by bone biopsy (Rabinowitz et al., 1976), the same subjects appeared to have much more rapid trabecular and cortical bone turnover than predicted by their models. The clear conclusion, if both the kinetic model and the biopsy results are accepted as correct, is that either the iliac bone is not part of the stable pool or that there must be exceptionally stable portions of the bone pool that outweigh the relative lability of iliac bone. In either case, it is clear that bone is not a single homogeneous pool nor is either major type of bone (cortical or trabecular) likely to represent a homogeneous pool of lead, this obviously having some implications for utilizing invasive and noninvasive methods of sampling of bone so as to determine its lead content.

20.7 BIOMARKERS An evaluation of potentially useful biomarkers of lead exposure requires consideration of a variety of specific issues. In common with all toxic exposures is the basic dose–response issue. In any such system, if dose–response characteristics are typical, well defined, and predictable, it makes little difference whether one focuses on measuring the dose or prefers to focus on a response variable. In the case of a toxicant such as lead, which has many diverse effects, it is usually necessary to define the response variable(s) of most importance to the investigator. In the case of mammalian (particularly human) studies, it is often useful to try to define the “critical” organ, tissue, or system. This definition presupposes that it is indeed possible to define the most sensitive or most important effect of the toxic agent. In the case of lead toxicity, especially childhood lead toxicity, the nervous system has been most commonly identified as being the “critical” organ or organ system. To the extent that the nervous system is accepted as the critical organ system, then any indicator of lead exposure/dose, whether measurement of lead per se in one or another tissue component or of some closely varying biological response, should closely define or reflect the extent of nervous system exposure to lead. Ideally, any response variable(s) used as the lead exposure/dose marker (s) should be based on nervous system toxicity or, at least, strongly correlate with nervous system toxicity. An important correlate of variation in response to a toxic agent such as lead is that it is often more difficult to detect a response in an individual than it is for groups of individuals. Hence, some particular measures of dose and/or response that may be useful in epidemiological studies may not necessarily aid much in the diagnostic categorization of the individual subject. An in-depth analysis of analytical methods, important considerations in selection of lead biomarkers, and interpretation of data for

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sensitive populations (e.g., maternal/fetal pairs, women of childbearing age, infants, and young children) was published by the National Research Council’s Committee on Measuring Lead in Critical Populations (1993). Lead exposure occurs via multiple routes, at variable external dose rates, and results in variable absorption depending on route(s) of absorption and other factors described elsewhere in this chapter. It is, therefore, generally very difficult to directly specify the “dose” of lead to which a subject is (or has been) exposed. The exception to this has been studies using stable radioactive lead tracers, wherein the absorbed dose can be quite carefully calculated by isotope dilution methods (Rabinowitz et al., 1976) or through careful measurement of diverse external sources (e.g., see Gulson et al., 1996). With the exception of such tracer methods, approaches to characterizing lead dosage in living subjects have generally been limited to measurement of lead in relatively easily obtained biological samples. All such methods encounter difficulties in sample contamination and analytical technique because of the very small amounts of lead typically found in the samples. Measurement of lead in whole blood has been most widely used, but has been criticized on practical and theoretical grounds. Difficulties potentially arising from contamination or due to technical aspects of measurement have largely been averted when blood lead analyses are done by a competent laboratory, and by the late 1990s, many laboratories could accurately detect blood lead concentrations less than 1 mg/dL. On the other hand, lead in blood represents only a small fraction of the total lead body burden and there is extensive turnover of lead in the body. Thus, the blood lead concentration at any given point of time can be considered to reflect both current and longer term past lead exposure (Mushak, 1989; Gulson et al., 1995, 1997, 1998a, 1998b, 1998c; Smith et al., 1996). It has been posited that lead in blood plasma is a better measure of lead available for internal transport to target tissues (e.g., neural tissues) and that such lead may, therefore, be a preferable measure of lead dose. However, it has been very hard to accomplish measurement of lead in plasma free of contamination, mainly due to the destruction of red cells in preparation of plasma samples. Because more than 95% of lead in blood is in red cells (DeSilva, 1981), even a small degree of contamination by red cell material could markedly alter plasma lead results. Urine lead reflects lead in blood in the sense that their stable lead isotope profiles are highly correlated, but lead concentrations in these two biological fluids are typically only weakly correlated (Gulson et al., 1998b). Thus, urine lead concentrations cannot serve to predict blood lead concentrations, particularly at exposures associated with blood lead values 510 mg/dL. Measurements of lead in urine may pose significant problems for a variety of reasons. First, as noted elsewhere in this chapter, urine lead has a complex relationship with lead dose. Also, the concentration of urine itself is affected by fluid intake. Thus, the variability of urinary lead excretion during the day has been found to be rather problematic when attempting to express lead dose (Gulson et al., 1998b). Variations in the concentration of urine itself (related to fluid intake) and the marked variability of urine lead excretion throughout the day mandate careful quantitative urine collections, which can be quite difficult to accomplish for children and experimental animals. Yet another problem is the contamination of urine samples with feces or other body products. A variant of urine lead measurement, often used diagnostically but infrequently in research studies, is the measurement of the amount of lead in urine after administration of a chelating agent. Such measurements presumably sample a larger pool of lead than unstimulated urinary lead excretion and have been held to define the “chelatable” pool of

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lead that should be available for further therapeutic chelation therapy. However, the “chelatable” pool varies with the agent used, and it is not particularly clear, in any real sense, as to what the “chelatable” pool represents biologically or how it can be directly related to response measurements. Because, with few exceptions, it is not routinely possible to directly sample body tissues other than blood or urine in living subjects, the measurement of lead in samples of other tissues has generally seen only very limited use. For example, since lead in the dentine of teeth represents a stable pool of lead reflecting cumulative past lead exposure, the measurement of lead in dentine of shed deciduous teeth has been used for some research on the effects of early lead exposure in children (Needleman et al., 1979). Also, bone biopsies have occasionally been used to assess quantities of lead in the more stable and larger pools of lead in the body (Rabinowitz et al., 1976; Aufderheide and Wittmers, 1992). Lastly, although sporadic efforts have also been made to use samples such as hair, nails, or saliva to assess lead exposure or dose in living subjects, all have been found to be unsatisfactory for general use for one reason or another and, at best, have only very limited applicability. Most direct measurements of lead in internal organs and tissues have been done on autopsy subjects (e.g., Barry, 1975, 1981; Drasch et al., 1974, 1997; Drasch et al., 1987; Drasch and Ott, 1988). However, the development of methods for noninvasive measurement of lead in tissues of living subjects has begun to emerge as a potentially viable alternative means for assessing lead levels in internal organs. In particular, the development of approaches to measurement of lead in bone in living subjects has progressed notably during the last decade or so (Nordberg et al., 1991). The human skeleton contains the great majority of body lead burden. The inactivity of the skeletal lead deposits was thought to reflect a very long half-time of lead in bone, and it was generally assumed that bone was homogeneous as a lead compartment and that the very long half-time would greatly delay transfer of lead from bone back to other tissues. Based on data from stable isotope studies, this is no longer a defensible concept. Current evidence indicates that bone is not only a set of compartments for lead deposition but is also a target of lead toxicity itself. Human bone appears to have at least two kinetic compartments for lead. Trabecular (spongy) bone lead is more mobile than lead stored in long, dense, or cortical bone (Skerfving, 1998). Also chelatable lead is well correlated with trabecular but less so with cortical bone (Schutz et al., 1987a, 1987b). In adults, long- term tissue (presumably largely bone) stores of lead contributed between 50 and 75% of the lead present in blood (Gulson et al., 1995, 1997; Smith et al., 1996). Young children, due to constant skeletal turnover during physiological remodeling processes that accompany somatic growth, recycle lead between bone and other tissue compartments. Rosen et al. (1989) reported that cortical bone (e.g., tibia) lead is correlated with and predictive of chelatable lead. During the past few decades, X-ray fluorescence methods have been developed to measure lead in bone noninvasively (Committee on Measuring Lead in Critical Populations,1993). Two general groups of XRF techniques can be distinguished based on their sampling of the fluorescence emitted either by k-shell or by l-shell electrons following radiation from an X-ray machine or other radiation source. Analyses of dosimetry, volume sampled, and precision for these instruments have been provided in the National Research Council’s report “Measuring Lead Exposure in Infants, Children, and Other Sensitive Populations” (1993). Such XRF methods have gained some use in a few epidemiological studies (e.g., see Hu, 1998), having been applied most successfully to groups with high lead exposures, for example, for persons living in high lead exposure environments, those occupationally exposed, or overtly lead-poisoned children. One concern is that the

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quantitation limits and precision of the XRF instruments may be too high for use in a general population (Rosen and Pounds, 1998). The rate of improvement in quantitation limits and precision of these instruments appears slower than the rate of declining levels of bone burden of lead in many countries. A wide range of methods have been employed, with varying success, to define the biological effects of lead. Such methods range from in vitro biochemical testing to observation and measurement of behavioral attributes of exposed subjects. Among the widely diverse array of biological effects shown to be caused by lead, the impairment of heme synthesis has long been recognized as being a key class of pathophysiological effects highly responsive to variations in lead exposure/dose. With the reliable detection of lead-induced impacts at several points along the heme synthesis pathway in red blood cells being facilitated by sampling of readily available blood in living subjects, certain indicators of lead-induced impairment at salient points in the heme synthesis pathway were used for many years as biochemical measures (i.e., biomarkers) of lead exposure and/or toxicity. Particular emphasis has been placed on the inhibition of two enzymes in the heme biosynthetic pathway, porphobilinogen synthetase and heme synthetase (Sassa et al., 1973; Piomelli, 1973). Porphobilinogen synthetase has been measured by activity levels and characterized by electrophoresis as to its phenotypic variability (Doss et al., 1982). Inhibition of heme synthetase, as a mitochondrial bound and dependent enzyme, has been gauged primarily by the accumulation of its porphyrin precursors, especially photoporphyrin IX. Accumulation of zinc protoporphyrin is strongly and logarithmically correlated with blood lead concentrations in both children (e.g., Piomelli et al., 1973, 1982; Roels et al., 1976) and adults (e.g., Grandjean and Lintrup, 1978; Lilis et al., 1978). Among children, the threshold for response is thought to fall within a blood lead concentration range of 15–20 mg/dL whole blood (Piomelli et al., 1982; Hammond et al., 1985). Analysis of free erythrocyte protoporphyrin (FEP) had previously been used in screening children to identify lead poisoning. However, because accumulation of protoporphyrin is also seen with iron deficiency, increased FEP is not a change specific only to lead, and because iron deficiency is sufficiently common among lead-exposed populations, the measurement of erythrocyte protoporphyrin (EP) poorly distinguishes between iron deficiency and lead excess (Mahaffey and Annest, 1986). This as well as the threshold for EP change being higher than for blood lead levels that are now recognized to be of concern based on neurobehavioral changes has led to curtailing of use of erythrocyte protoporphyrins as a lead exposure screening method. Another area of biochemical investigation, that is, the evaluation of levels of neurochemical mediators in blood and urine, has also been of some interest, but has thus far produced often conflicting results in the hands of different investigators. It appears clear that in consideration of all areas of biochemical investigation, such markers have not produced, to date, reliable and valid measures of response to lead exposure, at least in regard to neurological effects of lead. It remains to be seen if any other types of lead-induced changes in signature indicators of lead-related neurotoxicity, especially any of those resulting from decreasingly lower lead exposure levels, may emerge as useful biomarkers in the future. One possibility might be the use of certain electrophysiological changes, for example, altered brainstem auditory evoked potentials (BAEPs), that are discussed in the next section. Or, perhaps, further advances in the pioneering use of magnetic resonance imaging (MRI) or of magnetic resonance spectroscopy (MRS) methods to detect lead-induced neurotoxic effects (also discussed below) may ultimately generate new results that may prove to be useful in deriving new biomarkers for lead.

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20.8 HEALTH EFFECTS As noted earlier, it has long been known that high-level lead exposures can cause quite serious health effects in human children and adults. Such effects include severe neurological, renal, and hematological impairments that typify classically defined lead ‘poisoning” or “intoxication” that continued to be of much medical concern well into the latter part of the twentieth century. However, extensive research conducted especially since the 1960s–1970s has led to the recognition that lead exerts notable impacts on many different tissues and organ systems, including some impacts seen at very low exposure levels extending down to only slightly above current U.S. population means. This has been accompanied by parallel shifts in public health protection/medical attention to focus on reducing human lead exposures as the primary approach to dealing with lead as a continuing important public health issue. Reflecting this widespread shift of public health/medical interest in many countries, key emphasis is accordingly placed here on the discussion of health effects induced by low-level lead exposures. Unfortunately, the present space limitations preclude comprehensive discussion here of the full range of the rapidly expanded new information on lead-induced health impacts. Thus, rather than attempting to summarize the entire broad scope of newly documented lead effects here, the ensuing discussion focuses most strongly on neurotoxic impacts that, as an overall class, have come to be seen as key “signature” or “critical” effects of low-level lead exposures. A few examples of newly demonstrated low-level lead exposure impacts on other organ systems (e.g., renal, cardiovascular, immune) are also highlighted, and the reader is referred to other reviews for much fuller discussions of lead impacts on various organ systems and functions. 20.8.1

Neurotoxic Effects of Lead

Among the best known and most widely recognized “subclinical” impacts of low-level lead exposures are decrements in IQ and impacts on other global measures of neurocognitive abilities. Following pioneering work in the 1970s, for example, by Needleman et al. (1979), numerous epidemiological studies during the 1980s and 1990s evaluated the lead effects on the higher order integrated neurological functions (as indexed by lead impacts on intelligence measures, perceptual-motor coordination, and various other neurobehavioral end points). As noted in the prior version of this chapter, such studies (e.g., see Bellinger et al., 1986; Baghurst et al., 1992; Dietrich et al., 1990, 1993; Wasserman et al., 1997) substantiated that lead has adverse neurocognitive effects at very low levels of exposure, especially of the fetus and infant, as demonstrated by investigations in multiple cultures. Several major prospective, longitudinal epidemiological studies were highlighted as having shown impaired intellectual functioning in childhood following increases in blood lead across a range of approximately 10–30 mg/dL whole blood, even after control for social and demographic conditions associated both with exposures to lead and with lowered (slowed) development scores (Bellinger et al., 1986; Dietrich et al., 1993a, 1993b; Baghurst et al., 1992; Wasserman et al., 1997). These studies found an approximate 4–6-point decrease in subsequent IQ (when measured at about age of 6 or 7 years, but not earlier). Of interest, analogous to some earlier reports by Baghurst et al. (1995) and Dietrich et al. (1993) of lead impairment of visualmotor integration at this range of exposures, lead impacts on perceptual-motor skills were also found by Wasserman et al. (1997) and were suggested as possibly being more sensitive to lead exposure than language-related aspects of intelligence. Despite the demonstration of neurocognitive effects across geographic areas, social class, and cultures by the above

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studies, it should be noted that such findings were not obtained by all other analogous contemporary studies. More recent studies have continued to find evidence of impaired cognitive abilities being associated with low-level lead exposure of children, even at still lower levels than those previously thought to be harmful. Extensive, detailed reviews of such studies and those of other types of neurotoxic effects of lead in children and adults can be found in the recent U.S. EPA (2006) lead assessment and in the overview of neurotoxic effects of lead in humans by Bellinger (2008). For example, as discussed in U.S. EPA (2006), in the largest available new cross-sectional study, Lanphear et al. (2000) found relationships between the blood lead concentrations and the cognitive deficits in a nationally representative sample of 4853 U.S. children (all NHANES III participants), aged 6–16 years (having a geometric mean blood lead value of 1.9 mg/dL, with 97.9% being below 10 mg/dL). In multivariate analyses, significant covariate-adjusted associations were found between the blood lead levels and the two subtests for visual-motor skills and for short-term and working memory for all children and for those with blood lead 510 mg/dL, as well as with the visual-motor subtest for children with blood lead levels 57.5 mg/dL. U.S. EPA (2006) also noted that numerous other recent longitudinal studies have consistently observed effects on IQ in children at blood lead levels 510 mg/dL. Perhaps of most import, a large international pooled analysis of 1333 children from 7 different cohorts (by Lanphear et al., 2005) was highlighted as estimating a 6.2 point decline in full-scale IQ per increase in blood lead concentration from 1 to 10 mg/dL. Also of much importance is an observation across several of the new studies of a nonlinear relationship between blood lead concentrations and IQ or other neurobehavioral outcomes, with larger impacts being seen per unit increase in blood lead levels below 10 mg/dL than above that level. This otherwise nonintuitive dose–response relationship may be plausible, as noted by U.S. EPA (2006), if different underlying biological mechanisms (e.g., early CNS neurodevelopmental processes) are initially affected at relatively lower lead exposures than are other processes that may be disrupted in producing classic indicators of frank lead poisoning, with the dominant mechanisms at low exposure levels perhaps being very rapidly saturated versus less rapidly saturated mechanisms becoming predominant at higher exposure levels. However, it should be noted that although the reported nonlinear relationship between lead effects and neurocognitive functions appears to be gaining wide acceptance, some associated controversy remains, as reflected by the note by Bowers and Beck (2006) and the ensuing series of comments published in Neurotoxicoloy, Vol. 27, 2006, and Vol. 28, 2007. As stated in the prior version of this chapter, other neurocognitive changes and long-term educational, behavioral, and social consequences of low-level lead exposure have also been identified. For example, Bellinger et al. (1992) reported a higher rate of retention in grade and other results reflecting learning difficulties among higher blood lead children. Several more recent studies, as summarized in U.S. EPA (2006), have since confirmed analogous low-level lead exposure impacts on academic achievement. In one study highlighted by U.S. EPA, Lanphear et al. (2000) used multiple linear regression analyses of standardized academic achievement measures for the 4853 NHANES III children noted above as aged 6–16 years (mean blood lead, 1.9 mg/dL) and found blood lead to be significantly related to decrements in both reading and arithmetic achievement scores. Such decrements were also found in analyses stratified by blood lead concentrations to be inversely related to blood lead for those children with concurrent blood lead values 55 mg/dL. In yet another study noted by U.S. EPA (2006), from among 533 girls aged 6–12 years in Riyadh, Saudia Arabia (having a mean blood lead of 8.1 mg/dL), percentile of class rank was significantly associated with blood lead

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level in a subset of those with blood lead levels 510 mg/dL (Al-Saleh et al., 2001). Also, significant associations were noted as being found between blood lead values (mean of 11.4 mg/dL) and poorer math and vocabulary scores achieved by 594 second graders in Mexico, with segmented regression analyses showing the slope for the lead effect to be significantly steeper for blood lead values below 10 mg/dL (Tellez-Rojo et al., 2006). The effects on academic achievement seen in the above studies were statistically significant even after adjustment for IQ, thus raising the possibility (as posed by U.S. EPA, 2006) that the impairment of neurocognitive functions besides those indexed by global intelligence measures may contribute to lead-induced impacts on learning and academic achievement. It should also be noted, as per U.S. EPA (2006), that academic achievement decrements seen in the above studies may possibly be attributable to earlier (but unmeasured) higher pediatric blood lead levels (that usually peak before 3 years of age, then decline). As for lead impacts on neurobehavioral end points besides the above global measures of intelligence or academic achievement, U.S. EPA (2006) also noted that epidemiological studies have evaluated lead effects on more specific cognitive abilities, for example, attention, memory, visuo-spatial processing, and executive functions (impulse control, planning, other integration of higher order cognitive processes, etc.). Recent such studies were further noted as having shown relationships between blood lead and impacts on attentional behaviors and executive function among cohorts of children (varying in age range from 4–5 years to 19–20 years), even in those cohorts with more than 80% of subjects having concurrent blood lead values 510 mg/dL. Epidemiological studies were further noted as having demonstrated childhood lead exposure to be associated with disruptive/antisocial behavior, with such effects apparently persisting into adolescence and early adulthood. As an example, in following from ages 7–11 years, the same cohort of children shown by Bellinger et al. (1992) to have lead-related increased retention in grade, Needleman et al. (1996) found that lead exposure also increased risk for antisocial and delinquent behavior at 11 years of age. Dietrich et al. (2001) also reported behavioral disturbance and/or delinquency among young adults to be significantly related to blood lead measures obtained for them at various earlier time points (prenatally, at intervals during infancy and childhood, etc.) during participation in the Cincinnati prospective cohort lead study. Analogous long-term effects were reported by Burns et al. (1999), who observed that increasing blood lead across the range of 10–30 mg/dL adversely affected the behavioral and emotional development of children in the Port Pirie, Australia, cohort when evaluated at ages 11–13 years. In the same children, increasing blood lead concentration across the range of 10–20 mg/dL was also found to be associated at such ages with a three-point decline in mean IQ (Tong et al., 1996). Lastly, Needleman et al. (2002) found bone lead to be one of the strongest predictors of adjudicated delinquency among high-school-aged White and African-American subjects living in the Pittsburgh, PA, area. Specific neurological substrates and biochemical mechanisms that may be perturbed by lead in contributing to increased behavioral disturbance/ antisocial behavior remain to be more definitively characterized. However, U.S. EPA (2006) indicated that Lidsky and Schneider (2003) have noted that lead affects numerous brain sites and processes involved in impulse control, and Needleman et al. (2002) proposed that lead impacts on cognitive function and academic performance may indirectly contribute to antisocial behavior/delinquency. In addition to the above types of effects on intelligence and other higher order integrative functions, lead has also been shown to exert notable effects on more basic levels of sensory, motor, and sensory-motor integrative functions in children. For example, as noted by U.S. EPA (2006), epidemiological studies have shown lead effects on hearing thresholds and

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other features of auditory processing in children that appear to persist into young adulthood, including(a) observation by Schwartz and Otto (1987) of significant elevations in pure-tone hearing thresholds (at frequencies within the range of human speech) among more than 4500 NHANES II participants (aged 14–19 years); (b) replication by Schwartz and Otto (1991) of such findings in a sample of approximately 3000 subjects (6–19 years old) in the Hispanic Health and Nutrition Examination Survey (HHANES), even at blood lead levels below 10 mg/dL; (c) observation by Dietrich et al. (1992) of associations between higher prenatal, neonatal, and later postnatal blood lead concentrations in 215 children from the Cincinnati Lead Study and poorer scores on a test of central auditory processing (SCAN) at age 5 years; and (d) demonstration by Osman et al. (1999) of significant associations between concurrent blood lead levels and increased hearing thresholds among 155 children (aged 4–14 years) in Poland, which remained significant in analyses restricted to data for blood lead levels 510 mg/dL. As also noted by U.S. EPA (2006), Bellinger (1995) has suggested that such lead-induced impacts on hearing and auditory processing may be a mechanism contributing to learning impairment by lead. The above studies indicate that such sensory effects occur at rather low exposure levels (e.g., blood lead concentrations 510 mg/dL), although it should be noted again that some of the observed auditory effects could, potentially, derive from somewhat higher earlier peak blood lead concentrations. Several epidemiological studies have also characterized lead-related neuromotor deficits at relatively low levels of exposure, as discussed in U.S. EPA (2006). Dietrich et al. (1993a , 1993b), for example, reported that both pre-and postnatal blood lead concentrations were significantly related to poorer scores on tests of bilateral coordination, visual-motor control, upper limb speed and dexterity, fine motor control, and postural stability among Cincinnati Lead Study children at 6 years of age, with strongest associations being seen with concurrent lead levels (mean of 10.1 mg/dL). Later, 78-month postnatal blood lead levels were associated with poorer fine motor skills at the age of 16 years. Other studies were also noted as finding(a) associations between lifetime average blood lead levels through 54 months of age and poorer fine motor and visual motor function among Yugoslav children (Wasserman et al., 2000a), and (b) significant associations between blood lead values (mean 5.0 mg/dL) and increased reaction time, postural sway oscillations, and action tremor among 110 preschool Inuit children in Canada, even when data from the 10% of the children having blood lead levels 410 mg/dL were excluded from analyses. In addition to the above types of impacts, lead has also been shown to affect various electrophysiological measures of sensory and motor neurological responses. The prior edition of this chapter noted that peripheral nerve conduction (Seppalainen et al., 1979), as well as visual (Araki et al., 1987) and auditory (Schwartz and Otto, 1987) brainstem responses, was altered at relatively low lead exposure levels. Also highlighted were other epidemiological findings of significant lead-related impacts on visual-evoked potential interpeak latencies being seen (Altmann et al., 1998) in an environmentally exposed population of 6-year-old children, with mean blood lead of 4.2 mg/dL and 95th percentile value of 8.9 mg/dL (Walkowiak et al., 1998). During the past 10 years or so, certain innovative new approaches have begun to be used to evaluate neural substrates and neurochemical processes potentially perturbed by lead and possibly contributing to neurobehavioral impacts of lead. For example, U.S. EPA (2006) noted that MRI and MRS have come to be used to evaluate lead-exposed children, with several studies comparing subjects with elevated blood lead levels (420 mg/dL) to control subjects (blood lead 510 mg/dL). It was highlighted that although all subjects had normal MRI, the elevated lead subjects showed significant reductions in the ratios of

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N-acetylaspartate to creatine and phosphocreatine in frontal gray matter relative to the lower lead control subjects (Trope et al., 2001) and reduced peak values of choline, creatine, and N-acetylaspartate were seen in all the four brain regions of the lead-exposed children (Meng et al., 2005). One or the other such lead effects, it was suggested, could be related to lower neuronal density/neuronal loss, decreased cell membrane turnover, myelin alterations (possibly leading to CNS hypertrophy), or less neuronal cell viability. Also, in another study using functional MRI methods to follow a subsample of 48 young adults (aged 20–23 years) from the Cincinnati Lead Study, higher childhood average blood lead levels were found to be associated with reduced activation in Broca’s area (a brain area involved in speech production) while performing an integrated verb-generating/finger tapping task (Cecil et al., 2005; Yuan et al., 2006). The prior version of this chapter also noted that (a) the acute neurological damage in adults due to high levels of inorganic lead exposure and neurological consequences of exposure to organic lead compounds have been recognized for decades, and (b) additional findings on adverse neurologic effects of lead, including at exposure levels not previously recognized as harmful to the adult, have been elucidated within the past 10 years. For example, Hanninen et al. (1998) reported that in the studies of workers whose blood lead levels had never exceeded 2.4 mmol/L (50 mg/dL), lead was found to be associated with decrements in visual-spatial and visual-motor functions, verbal comprehension, and attention, as well as increased symptoms of impaired well-being as rated by psychological assessments of mood, whereas blood lead increases approximately from 2.4 mmol/L (50 mg/ dL) to 4.9 mmol/L (100 mg/dL) caused persisting, possibly permanent impairment of CNS function (Hanninen et al., 1998). 20.8.2

Other Effects

The previous edition of this chapter noted that adverse renal system effects typically have been described in many earlier reviews of high lead exposure effects. It further noted that acute effects of exposure to high concentrations of lead result in proximal tubule damage manifested by glycosuria and amino acid uria and that overt nephropathy appears to develop when blood lead levels exceed a threshold of 60 mg/dL (about 2.9 mmol/L), as reviewed by Loghman-Adham (1997). However, early renal tubule dysfunction secondary to far lower levels of lead exposure was also indicated as having been detected by measuring of urinary excretion of low molecular weight proteins (81 and 92 microglobulins, retinol binding protein) or the lysosomal enzyme NAG, as well as other brush border proteins (Loghman-Adham, 1997). A cross-sectional study by Fels et al. (1998) was further highlighted as comparing changes in urinary or serum markers of function or integrity of specific nephron segments in children, with those having a mean blood lead concentration of 13 mg/dL showing increased excretion rates for prostaglandins and thromboxane B2, epidermal growth factor, 92-microglobulin, and Clara cell protein compared to children with mean blood lead levels averaging less than 4 mg/dL. This overall pattern of glomerular, proximal, and distal tubular, and interstitial markers was noted as being similar to that earlier found among adults, but occurring at lower blood lead concentrations than in adults (Fels et al., 1998). The clinical significance of these changes in excretion of low molecular weight proteins and/or lysosomal enzymes was noted, however, as not yet being fully clear. Additional evidence derived from several U.S. and European general population studies during the past 2 decades substantiates further that low-level lead exposures affect the above

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or other indicators of altered renal function, as discussed in recent reviews/assessments (Gonick, 2002; U.S. EPA, 2006). For example, U.S. EPA (2006) noted that the Belgian Cadmibel study(a) was the first large environmental study to adjust for multiple renal risk factors; (b) evaluated several renal outcome measures among a general adult European population; and (c) found decreased creatinine clearance to be related to blood lead and zinc protoporphyrin concentrations in both men (blood lead mean of 11.4 mg/dL and range of 2.3–72.5) and women (blood lead mean of 7.5 mg/dL and range of 1.7–60.3 mg/dL), thus raising concerns that the exposure/dose threshold for adverse lead effects on renal function among the general population might be much lowered than earlier thought based on studies of occupationally exposed workers. U.S. EPA (2006) also highlighted several other published analyses (Payton et al., 1994; Kim et al., 1996; Wu et al., 2003; Tsaih et al., 2004) of data from the Normative Aging Study (a long-term study of Boston area adult men, aged 21–80 years, with participants initially having been recruited in the 1960s and undergoing periodic follow-up evaluations), which demonstrated relationships between blood and/or bone lead concentrations and various indicators of reductions in measured or estimated creatinine clearance, including among men with peak blood lead levels below 10 mg/dL in some analyses. Also highlighted by U.S. EPA (2006) were the results of analyses by Muntner et al. (2003) of associations between blood lead and renal outcomes among 15,000 NHANES III adult participants during 1988–1994. The analyses were stratified because of an interaction between blood lead and the hypertension (HTN) (as per leadrelated cardiovascular effects noted later), with (a) mean blood lead concentrations being 4.2 mg/dL in hypertensive and 3.3 mg/dL in normotensive subjects; (b) the prevalence of elevated serum creatinine levels (indicative of reduced renal clearance of creatinine) being about 10 times higher in hypertensive (11.8%) than in normotensive (1.8%) subjects; and (c) higher blood lead levels being associated with a higher prevalence of chronic kidney disease in diabetics among the nonhypertensive subjects. Lastly, U.S. EPA (2006) also noted ˚´ kesson, 2006) to be significantly related ˚´ kesson et al., 2005; A that blood lead was reported (A to indications of altered renal function (e.g., reductions in estimated creatinine clearance) among 820 women, aged 53–64 years, in the Lund area of Sweden, with the association being apparent over the entire blood lead range (mean blood lead of 2.2 mg/dL). U.S. EPA (2006) also discussed an interesting quantitative comparison (see Chapter 6, Figure 6.8 in the U.S. EPA document) of estimated impacts of lead exposure on renal creatinine clearance derived by the various analyses discussed above and noted that the slopes from those studies fell in the range of 0.2 to minus 1.8 mL/min change in creatinine clearance per 1 mg/dL blood lead increase. Overall, the above analyses appear to substantiate well that the impairment of renal function by lead occurs at much lower exposures/doses than those previously thought to be harmful for nonoccupationally exposed adults in the general population. This appears to include lead exposures resulting in blood lead levels extending to well below 10 mg/dL, but one cannot completely rule out the possible attribution of effects observed late in adulthood in some studies to somewhat higher (but still likely rather low) unmeasured peak exposure levels that may have occurred earlier in the study participants’ lives. As noted at the outset of this section, lead has now come to be recognized as exerting notable effects, including at relatively low exposure levels, on a number of organ systems besides those (neurologic, renal, etc.) classically recognized a key target organs for the metal. Among the most important classes of newly demonstrated types of effects are cardiovascular impacts of lead exposure. U.S. EPA (2006) assesses key information on leadrelated cardiovascular effects that has emerged during the last several decades. U.S. EPA

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(2006) review noted that epidemiological studies that have examined relationships between human blood lead levels and blood pressure have generally found positive associations, even after controlling for confounding factors such as tobacco smoking, exercise, body weight, achohol consumption, and socioeconomic status. It further highlights recent meta-analyses of such studies that found robust, statistically significant, though small effect size, associations between blood lead concentrations and blood pressure. The Nawrot et al. (2002) metaanalysis was cited as an example in finding that a doubling of blood lead corresponded to a 1 mmHg increase in systolic blood pressure, and it was noted that while not necessarily being clinically meaningful for a given individual, a population shift in blood pressure of 1 mmHg is likely of concern in terms of associated increased risk for more serious cardiovascular outcomes (heart attack, cerebrovascular events, etc.). U.S. EPA (2006) further noted that the majority of recent studies evaluating relationships between the bone lead and the cardiovascular effects found strong associations between the long-term lead exposure (indexed by bone lead concentrations) and the arterial blood pressure. Also noted, too, was highly supportive evidence from numerous animal studies showing that low-level lead exposures for extended periods of time result in the eventual onset of arterial hypertension, which persists long after exposure cessation, with both in vivo and in vitro toxicology studies (Gonick et al., 1997; Vaziri et al., 1997) providing strong evidence that oxidative stress, at least in part, plays a key role in mediating lead-related HTN. An excellent review of pertinent literature on this subject has been provided by Vaziri (2002). Impacts on the immune system have also emerged during the past 20–30 years as yet another important class of lead-related toxic effects subsumed as part of a much broader array beyond those earlier classically defined as typifying lead toxicity, again including notable lead impacts seen with rather low-level exposures. Such lead-related immune system effects have recently been assessed by U.S. EPA (2006) and in a review by Dietert and Piepenbrink (2006). Very notably, as summarized well in U.S. EPA (2006), lead effects on nonhuman animal immune systems appear to include the targeting of T cells and macrophages, with lead-induced alterations being typified by (a) an increased inflammatory profile for macrophages (e.g., elevated tumor necrosis factor-alpha, oxygen radical and prostaglandin production) and (b) skewing of the T cell response away from T helper 1(Th1)-dependent functions toward T helper 2 (Th2)-dependent functions. Resulting impacts on immune system function, as noted by U.S. EPA (2006), include increased production of Th2 cytokines (e.g., IL-4, IL-10) and certain immunoglobulins (e.g., IgE), decreases in Th1-associated cytokines (interferon gamma and IL-12) and Th1 functions (e.g., the delayed-type hypersensitivity, or DTH, response); but not major immune cell population changes (which complicates interpretation of human epidemiological study results). Also noted was an approximate order-of-magnitude age-related difference in immune system sensitivity to lead between the perinatal period and adulthood, with lead effects on immune function being seen at blood lead levels distinctly below 10 mg/dL following gestational or perinatal exposures. Also emphasized was a key point noted by Dietert and Piepenbrink (2006), that is, given that the most informative sources of functionally reactive immune cells (e.g., antigen-reactive ones in lymphoid organs and lymph nodes) are not readily accessible in humans, the circulating lymphocytes and serum or plasma immunoglobulin (e.g., IgE) levels must serve as readily accessible surrogate indices of immune status in human studies. Most importantly, it was highlighted that similar immune system effects have been observed in both humans and laboratory animals in terms of positive associations being seen between the circulating IgE levels and the bood lead concentrations following early life lead exposures, even at blood lead levels below 10 mg/dL.

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As for the other types of lead health effects, the prior edition of this chapter included brief discussion of reproductive impacts, noting that comparatively recent investigations of the reproductive effects of lead had most extensively focused on male reproductive toxicity (as reviewed by Apostoli et al., 1998). Perhaps of most note is the fact that based on review of 32 experimental animal studies, 22 human epidemiological studies, and one case report in humans, Apostoli et al. (1998) concluded that when blood lead concentrations exceed 40 mg/dL, there appears to be associated decreases in sperm count, volume, motility, and morphological alternations (with a possible effect also on endocrine profile), but dose– response relationships, particularly possible thresholds for effects, remained poorly understood. More recently, Bellinger (2003) and U.S. EPA (2006) have provided additional overviews of lead-related reproductive effects. Also of note is the fact that during the past 20–30 years or so, lead and lead compounds have been recognized as being carcinogenic to animals (IARC, 1987). The prior edition of this chapter also noted that better documentation of possible lead carcinogenicity among humans has begun to emerge, and cited a review of the carcinogenicity of lead by Vainio (1997) following publication of two cohort studies among smelter workers (by Lundstrom et al., 1997 and by Cocco et al., 1997). Overall, Vainio (1997) concluded that a long-term, highlevel exposure to lead compounds is associated with an increased risk of cancer and that the “weight of evidence is beginning to be convincing enough concerning kidney and even lung cancer” for humans. In addition, more recent extensive reviews have been published by both Silbergeld (2003) and U.S. EPA (2006) which evaluate available human and/or animal evidence concerning carcinogenicity effects of lead. The broad lead assessment by U.S. EPA (2006) covers the latest available information not only on all of the types of lead health effects discussed above, but also other classes of lead effects not addressed to any great extent here due to space limitations. It is recommended that the reader must also consult the discussion by Hu et al. (1998) of lead effects on bone and teeth.

20.9 MECHANISMS UNDERLYING LEAD TOXICITY The prior edition of this chapter noted that early investigations into lead toxicity focused on mechanisms underlying the relatively gross effects of lead observed at very high levels of exposure. For instance, the study of kidney pathology in lead-exposed animals and humans was noted as demonstrating deposition of lead in the cells of the proximal tubule, especially in nuclear material (Goyer and Rhyne, 1973), and such studies were also noted as documenting mitochondrial degeneration. Although these studies are consistent with the observed proximal tubular dysfunction seen in lead toxicity, it was also highlighted that they do not fully clarify the mechanism of this toxicity. Similarly, it was noted that studies of the heme biosynthetic pathway, though interesting, reveal certain inconsistencies that remain to be explained (Scott et al., 1971). Lead certainly inhibits several enzymes in this pathway, for example, porphobilinogen synthetase and heme synthetase in particular. However, of these, heme synthetase is of special interest, because it is a mitochondrial-bound enzyme, suggesting the possibility that lead inhibits this enzyme simply by altering mitochondrial function rather than by specifically affecting the enzyme per se. It was further noted that the mechanisms for nervous system toxicity as a major focus of attention remained unclear, but a number of important insights were highlighted. For example, the importance of lead effects on calcium-dependent systems appears to come

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as close as to an overarching explanation currently available. That is, ultimately, all lead toxicity may be related to calcium transport or to other mechanisms underlying alteration of functioning of calcium-dependent systems. Also, as noted elsewhere in this chapter, lead accumulates in the nervous system preferentially in areas very metabolically active and rich in mitochondria, and lead has been shown in vitro to be a potent inhibitor of mitochondrial function, possibly by competing with mitochondrial uptake of calcium. Lead compounds, it was noted, may have a variety of targets within the nervous system (Bondy, 1988), and before discussing what some of these targets might be, it may first be useful to classify broadly the types of lead compounds that are of environmental concern. As already pointed out, the inorganic lead compounds are generally of greater environmental concern. A brief discussion of properties of organolead compounds may help to clarify our knowledge about biochemical mechanisms of neurotoxicity for all classes of lead compounds. Consideration of the chemistry of organolead compounds suggests that they differ fundamentally in both chemical and biological properties from ionic compounds of the same metal (Grandjean and Grandjean, 1984). For example, it is well documented that the toxicity of tetraethyl lead results from the breakdown of the compound in the organism to a salt of triethyl lead. The triorganolead compounds form a very distinctive neurotoxic class. Their actions are probably brought about by two distinctive chemical properties: one depends on the lipophilic (rather amphoteric) nature of their chlorides and hydroxides and their affinity constants that allow dissociation at biological concentrations of hydroxyl and chloride ions; the other derives from the potentiality for five coordinate binding. Therefore, it is not surprising that the behavioral toxicity effects of alkyl leads do not closely resemble those of inorganic lead. Some similarities do exist that may be associated with the degradation of the alkyl lead to the stage of divalent lead in the organism. Even a small amount of metabolism of this type could become significant because of the much altered tissue distribution of lead as a result of the lipophilic properties of organolead compounds. This suggests that more attention should be paid to the possible complexes of inorganic lead with hydrophobic ligands such as hemic acid. The general lack of involvement/importance of the divalent lead ion in organolead compound toxicity is also supported by the observation that the usual chelators have little effect on intoxication from organolead compounds. The study of chemical mechanisms of inorganic lead compound toxicity is complicated by a number of factors. As already mentioned, the amount of lead absorbed after its oral administration can vary significantly depending on animal species, age, diet, and on both the chemical and physical form of the inorganic lead compound ingested. For example, in the absence of food in the gut (as during fasting), the lead compound can apparently be more readily acidified and solubilized for tissue uptake. Because of the rather ubiquitous occurrence of lead in the environment and its relatively high toxicity, one must also be concerned about lead contamination of the diet, and even of laboratory reagents (Simons, 1989) used in studies in vitro. Separation of the influence of nutritional status on biokinetics of lead from specific neurotoxic events is another problem that needs more attention. There are many ostensibly different toxic effects of lead described in the literature. It seems likely that there are only a few triggering events in biological systems that would account for the rather potent neurotoxic effects of lead. These in turn could result in a range of secondary effects. It follows also that the important initiating events will depend in some way on rather distinctive chemical properties of divalent lead, since all other divalent metals produce a toxic syndrome that is qualitatively or quantitatively different from that of lead. Considerations of lead chemistry should also permit one to rule out certain possible initiating

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biochemical pathways. For example, the relatively low affinity of lead for sulfur suggests that interaction with specific sulfur-containing proteins such as metallothionein would not be critical. Carboxyl-oxygen-containing amino acids are more likely points of interactions with proteins, and this possibility is supported by the findings of lead-containing nuclear inclusion bodies in tissues that are rich in acidic amino acids (Choie and Richter, 1972). Since divalent lead is a reasonably stable oxidation state, it seems likely that the divalent lead ion in some way mediates the range of effects seen and that metabolic redox pathways involving tetravalent lead are not important. The toxic effects of divalent lead may be the result of either physical or chemical change in the biochemical systems with which it interacts. Lead may damage membranes by bringing about change in the ultrastructure of cellular components or by initiating oxidative damage. Membrane damage does appear to be a significant factor in lead toxicity, and myelin-containing membranes seem to be especially sensitive. Membrane damage by peroxidative processes may involve change in calcium homeostasis. Effects of lead on energy production could be related to direct interaction with mitochondrial membranes, altering ion transport, or changes in calcium homeostasis within the cell. Various studies have shown that lead accumulates in mitochondria, and this is associated with the inhibition of oxidative phosphorylation. Endothelial cells in brain capillaries contain three–five times more mitochondria than do endothelial cells in the vessels of other organs (Oldendorf and Brown, 1975). This difference may reflect the high rate of metabolic activity necessary to maintain the active transport of ions across the blood– brain barrier and may explain the susceptibility of these cells to a wide variety of toxic compounds that cause brain edema. Mitochondria may be a critical subcellular target for the toxic effects of lead. Inorganic lead has also been shown to compete with some essential divalent metals at several different levels (Chisolm, 1980). These include, in particular, calcium, zinc, copper, and iron. These interactions can occur at several levels, including absorption from the gut, transport across the blood–brain barrier, and at the synapse. For example, active calcium uptake by mitochondria is a critical process required to maintain calcium at very low concentration in the cytosol of cells. Inhibition of calcium accumulation by mitochondria may involve a direct blockage of the calcium pumps, but this may also be attributed to depletion of ATP or key intermediates involved in its synthesis, such as inorganic phosphate. No single metal deficiency shows symptoms identical to those seen in lead exposure. The chemistry of lead suggests that a variety of metals are likely to show altered distribution in biological systems in the presence of lead, with some associated biochemical changes. Dietary supplements of various minerals and vitamins do not completely protect against the toxic effects of lead, suggesting also that there are more critical biochemical changes associated with the potent neurotoxic effects of lead. Other biochemical interactions that may be important in explaining the toxic responses of lead include the possible inhibitory effects of lead on the cholinergic system and activation of catecholaminergic function. This may be related to the calcium agonist property of lead. Lead may be equipotent with calcium in binding to calmodulin. Lead may also be involved in direct reactions carried out by mixed-function oxidases. Complex secondary interactions between organ systems are also a possible factor determining the overall pattern of toxicity. A possible interrelated sequence of events by which lead compounds could cause neurotoxicity, as expressed by behavioral change, has previously been suggested (Bondy, 1988). What aspects of the chemistry of lead may throw light on the important biochemical/molecular mechanisms of toxicity? The importance of the ligand exchange chemistry of divalent lead in the

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overall expression of toxicity has already been pointed out. The relative binding strengths of the various endogenous ligands for lead will determine the fate and distribution of lead in vivo and its competition with other metals such as calcium. Once lead is inside the cells, its eventual fate will be to bind to sites that are stronger than cytosolic pool chelators such as citrate. At slightly acidic pH, at which divalent lead ionic species are likely to be present, phosphate ligands are prime candidates because of the very large association binding constants involved here. The extreme tenacity for phosphate may be the most distinctive feature of lead chemistry relative to other divalent metal ions of the same basic types. Several types of phosphate ligands must be considered, including inorganic orthophosphates, ATP, and the phosphate groups of membrane lipophosphates and phosphorylated proteins. This primary mechanism underlying lead toxicity is compatible with the major reproducible biochemical findings described above, particularly the suggestion that mitochondria (where oxidative phosphorylation takes place) may be a target site in cells. It is of course also compatible with the ultimate localization of the lead in bone as insoluble phosphate salts. It was also noted that lead may function as a phosphate scavenger and siphon off minority phosphate species crucial to developing/proliferating cells, especially in neural tissues. If one of the functions of calcium is to store phosphate in the form of calcium phosphate, then lead would be an efficient antagonist for this process. In fact it is, in general, not clear if the effects of lead on calcium homeostasis are the result of direct competition between lead and calcium for binding sites and/or of their differential affinity for phosphate ligands, particularly inorganic phosphate. For example, the report of Markovac and Goldstein (1988) indicating that lead is a potent activator of protein kinase C could be interpreted as a direct effect of lead on sequences of protein–phosphorylating–dephosphorylating that do not involve the enzyme at all (a control experiment in the absence of enzyme was apparently not run). Note that phosphorylating–dephosphorylating sequences are of critical importance to energy transformation in cells and tissues, and this is particularly true of those in the nervous system during rapid growth and maturation. It was noted that the possible direct phosphate-deleting effects of lead resemble in many ways those seen in poisoning by nitrophenols (Clayton and Clayton, 1981), which are known to uncouple oxidative phosphorylation (which presumably also reduces the body’s reservoirs of high-energy phosphate compounds). Such uncoupling apparently stimulates oxidative metabolism and, in turn, heat production of the body. Oxygen consumption, body temperature, respiration, and heart rate are all increased. Some similar effects are also associated with the hyperthyroid state (Dratman, 1978). With regard to mechanisms of lead neurotoxicity, it is interesting to note that there is some evidence that hyperthyroidism is associated with reduced catecholamine production rates in both the peripheral nerve tissue and the brain. In studies on rat, McIntosh et al. (1989) reported that the effects of lead on catecholaminergic and cholinergic transmission are regionally specific within the brain, the midbrain, and the diencephalon, showing the greatest degree of change in concentrations of neurotransmitters (dopamine concentrations usually decreased) and activities of ratelimiting enzymes. See also Cory-Schlecta (1997) for information on relationships between lead effects on neurotransmitter systems and behavioral toxicity. Although lead neurotoxicity, poisoning by nitrophenols, and hyperthyroidism all involve different triggering mechanisms, these responses may have in common certain secondary effects related to the maintenance of cellular homeostasis and metabolism. The remarkable affinity of lead for phosphate may be the most sensitive primary event to explain the fact that the neurotoxic effects of lead are evident at low concentrations of lead exposure and after only a brief exposure period. This suggests that thyroid status and metabolic state might be

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important factors to consider in examining correlations between the measures of lead exposure and could be causally related to increased oxidative metabolism. It should be possible to test this hypothesis at the biochemical levels, perhaps in some cases using the tools now available in molecular biology. Since the above overview of hypotheses/information related to lead toxicity (especially lead neurotoxicity) appeared, reviews by U.S. EPA (2006) and White et al. (2007) have been published that include evaluation of further advances regarding mechanisms underlying lead neurotoxicity. The White et al. (2007) review, for example, highlighted four major areas of key advances in lead neurotoxicity during recent years: (1) experimental studies showing that stress markedly influences lead effects, possibly mediated by interactions of corticosteroid hormones with components of the mesocorticolimbic dopamine system of the brain (with heightened stress causing consequent elevations in circulating corticosteroid levels hypothesized to contribute to increased vulnerability to many diseases and other dysfuctions among lower socioeconomic status populations); (2) cellular models of learning and memory used to evaluate possible mechanisms of lead-related cognitive deficits (with studies of long-term potentiation in the rodent hippocampus having shown lead-related increased thresholds, decreases of magnitude, and shorter retention times for neural plasticity, and alterations in the form of adult neurogenesis in the hippocampus, which may contribute to learning impairment); (3) in vitro evidence for strong binding of lead to glucose-related protein (GRP-78), induction of GRP aggregation, and blocking of secretion of interleukin-6 (IL-6) by astroglial cells (findings that implicate lead in “chaperone deficiency” processes, which in the long-term could underlie protein confrontational diseases, e.g., Alzheimer’s disease); and (4) implication of lead exposure in the early development and in the later progression of amyloidogenesis in rodent brains during old age (thereby contributing to increased proteins associated with Alzheimer’s disease pathology). Overall, as noted by White et al. (2007), such findings provide compelling evidence for lead exposures having adverse effects on the nervous system, that environmental factors increase nervous system susceptibility to lead, and that lead exposures early in life may contribute to neurodegeneration later in old age.

20.10

TREATMENT OF LEAD TOXICITY

The prior edition of this chapter noted that metal chelation therapy has been used with some success to treat lead poisoning (Bondy, 1988). It further noted, however, that chelators used for this purpose can also remove essential elements resulting in kidney damage, and most of these drugs tend to have unpleasant side effects, that they are also most generally useful for acute rather than sustained therapy, and that the benefits derived are usually only transitory, since blood lead can be rapidly replaced from bone stores. Several other important points were made as well. For example, in view of the nature of lead chemistry and the possibility that lead is basically functioning as a phosphate scavenger, it is unlikely that a complexing/ chelating agent of the usual variety could be found that could compete effectively with phosphate in a ligand exchange reaction, but it may be possible to develop a derivative of phosphate that is sufficiently reactive and excretable to be useful in therapy. On the other hand, soluble forms of phosphate itself might offer some protection against toxic effects, perhaps until lead is deposited in bone. This approach does not reduce the body burden of lead, but it may buy time during the process of deposition of lead in bone, which at least offers transient sequestering of lead.

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It was further noted that the prospects for successful chemical treatment of long-term, low-dose lead toxicity are not promising and a number of pertinent points were articulated, as reiterated with only very minor change here. Chelating agents have been used mainly to treat the short-term, high-dose exposures. For children, CDC (1991) has recommended the use of chelation therapy only when blood lead concentrations exceed 45 mg/dL (42.17 mmol/L). It remains unclear whether chelation has sustained benefits for children with blood lead of 25–44 mg/dL. Criteria such as the persistence of elevated blood lead levels despite environmental intervention have also been offered as justifications for chelation (American Academy of Pediatrics, 1993). Deciding which children may respond to chelation therapy is complex. It was found that children with changes in erythrocyte protoporphyrin and hematological index are more likely to respond to chelators with markedly enhanced urinary excretion of lead. Concern has also been raised suggesting that at least some chelators (i.e., succimer) are more effective in removing lead from blood than lead from brain (Pappas et al., 1995; Cory-Slechta, 1988). Smith et al. (1998) cautioned with regard to basing judgments on changes in blood lead concentration on predictions of the impact of chelators on brain lead concentrations. The ratios of change in brain lead and change in blood lead differed over the duration of chelation therapy in rodents. The usefulness of nutritional therapy depends on the timing of its introduction, the severity of lead exposure, and underlying nutritional status. It is clear that marginal nutritional status is associated with increased prevalence of elevated blood lead concentrations. Data from national epidemiological surveys such as the National Health and Nutrition Examination Survey, conducted in the United States during the 1970s through the l990s, demonstrated that young children from socially disadvantaged, low-income, minority families are more likely to have a greater prevalence of elevated blood lead levels (Mahaffey et al., 1982; Pirkle et al., 1994) and of marginal nutritional status (Life Science Research Office, 1996; Mahaffey et al., 1986). The consumption of a higher calcium diet has also been shown to be inversely related to bone lead among women living in Mexico City (HernandezAvila et al., 1996). The physiological changes that accompany poor nutritional status for calcium and iron result in enhanced absorption of these required elements from the GI tract (Fullmer, 1997; Mahaffey, 1995). Because lead absorption also increases with these physiological changes, improving nutritional status results in reduced future absorption of lead. Some reports indicate reduced blood lead following treatment with iron (Granado et al., 1994). Short-term correction of low calcium intake has not been shown to alter blood lead, but it is clear that skeletal mineral can be mobilized for calcium under conditions of physiological stress and that lead will be released along with calcium during this mineral mobilization (Gulson et al., 1997a, 1998a). Whether increased calcium intake reduces this mobilization is a question still to be addressed. Among recommended “treatment” activities is identification through screening of cases for environmental/nutritional/pharmaceutical intervention. In 1991, the “case” definition was lowered to define childhood lead poisoning as a blood lead of greater than or equal to 10 mg/dL (0.48 mmol/L) (CDC, 1991; American Academy of Pediatrics, 1993). Although the 10 mg/dL action level still remain in effect, in a move away from universal screening (CDC, 1997), emphasis is now on geographic areas with higher prevalence of older housing (defined as where 27% or more housing was built before 1950) or presence of other risk factors (e.g., the child receives services from public assistance programs for the poor, or the child has a sibling or playmate who has had lead poisoning). Part of this change reflects the overall decline in blood lead concentrations in the United States.

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21 MERCURY Philippe Grandjean and Jesper B. Nielsen

21.1 INTRODUCTION Although the toxicity of mercury has been known since ancient times, its therapeutic effects were also utilized in a variety of drugs. In particular, mercurous chloride (calomel) was an important drug for syphilis treatment, although some patients inevitably became mercury poisoned. The occupational health risks were described by Bernardino Ramazzini 300 years ago. Risks due to environmental contamination came to the forefront in around 1960 when Minamata disease in Japan was found to be caused by mercury pollution from a local factory. Recent risk assessments include the Agency for Toxic Substances and Disease Registry (ATSDR, 1994) U.N. Environment Programme (UNEP, 2002) and the U.S. Environmental Protection Agency (2001); these sources may be consulted for further information and additional references.

21.2 CHEMISTRY Mercury exists in three oxidation states: Hg0 (metallic), Hgþ (mercurous), and Hg2þ (mercuric) mercury. In organometallic derivatives, mercuric mercury is covalently bound to one or two carbon atoms, and the organic part of the molecule is often an alkyl group or an alkoxialkyl group. The former compounds are more toxic because they are more easily absorbed and more slowly metabolized. In its elemental form, mercury is a dense, silvery-white, shiny metal, which is liquid at room temperature and boils at 357 C. At 20 C, the vapor pressure of the metal is 0.17 Pa (0.0013 mm Hg), and a saturated atmosphere at this temperature contains 14 mg Hg/m3, which is more than 100 times the occupational exposure limit. Mercury compounds differ greatly in their solubility. Thus, at 25 C, the solubility of metallic mercury, mercurous chloride, and mercuric chloride in water are 60 mg/L, 2 mg/L, and 69 g/L, respectively (IARC, 1994).


811

812

MERCURY

Exact data for the solubility of methyl mercuric chloride in water are lacking, but are known to be slightly higher than for mercurous chloride. Certain species of mercury are soluble in nonpolar solvents. These include elemental mercury and the halide compounds of alkylmercurials.

21.3 SOURCES Mercury is emitted to the atmosphere by “degassing” of the earth’s surface and by resuspension of mercury-containing particles previously deposited. Emissions from volcanoes and other natural sources are estimated to constitute about 1000 tons per year (UNEP, 2002). An additional annual emission of about 2500 tons comes from anthropogenic sources, the major source being energy production from fossil fuels, especially coals with high mercury contents (UNEP, 2002). Mercury is produced by the mining and smelting of cinnabar ore. It has been extensively used in chloralkali plants (producing chlorine and sodium hydroxide), but modern plant designs have now made large stores of mercury unnecessary. A myriad of mercurycontaining products have been in use, including compounds in paints as preservatives or pigments, in electrical switching equipment and batteries, in measuring and control equipment (thermometers and other medical equipment), in mercury vacuum instruments, as a catalyst in chemical processes, in mercury quartz and luminescent lamps, in the production and use of high explosives using mercury fulminate, in copper/silver amalgams in dental restoration materials, and as fungicides in agriculture (especially as seed dressings). Many of these uses are now being banned or phased out. According to data from the U.S. EPA, these efforts have been beneficial, as overall mercury emissions from industrial use in the United States have dropped 45% since 1990, and are still decreasing. One of the uses of liquid metallic mercury that has escalated during the last few decades is artisanal gold mining. Alluvial deposits of fine gold particles are often extracted using mercury. The gold particles are dissolved in the mercury as an amalgam, and the mercury is subsequently removed by heating with a gas torch. This use therefore exposes the gold miner to a substantial amount of mercury vapor, and also leads to extensive release of mercury into confined and sometimes ecologically sensitive areas. The annual consumption of mercury in such mining operations is about 650 tons, mainly in Asia, Central Africa, and Latin America. Another consequence of this practice is the contamination of soil, which can remain polluted for many decades. Some previous gold mining sites in the United States (e.g., Carson River, Nevada) are now recognized as being heavily contaminated with mercury, with estimated amounts of mercury residues exceeding 6000 tons. Deposition of sewage sludge and contamination from other industrial activities often involve mercuric salts with low solubility (i.e., sulfides). Ecological and human health implications inorganic mercury in the environment depend on the entent of mercury methylation. Recent studies using stable mercury isotopes (Harris et al., 2007) have documented that sedimentation of airborne mercury onto freshwater ecosystems within several months gets accumulated in fish in the form of methylmercury. Organomercury compounds now only find limited use as fungicides, but methylmercury was extensively used for this purpose in the past, until environmental effects were discovered. The less toxic methoxymethylmercury is still sometimes used for wood treatment or in the paper and pulp industry as an antislime agent. Thimerosal (ethylmercury salicylate) has been widely used as a preservative in the pharmaceutical industry, for example, in vaccines, but is now being phased out.

ENVIRONMENTAL EXPOSURES

813

The various uses of mercury and mercury compounds result in occupational exposures in a range of occupations, in most cases only involving mercury vapor and inorganic compounds. The industrial use of mercury may also lead to releases to the environment due to evaporation of releases in sewage water (IPCS, 1991). Localized problems relating to contamination of river systems and bays have been caused by contamination from chloralkali plants, paper and pulp industries, and pesticide factories. In Japan, the Minamata Bay became severely contaminated from a factory that used methylmercury as a catalyst in the production of acetaldehyde. In addition, airborne emissions from coalfired power plants and incinerators can cause contamination of lakes and rivers. Coal burning contributes 40% of the mercury emissions in the United States, and is associated with increased mercury deposition in Eastern Canada and New England (Rice and Hammitt, 2005). Elemental mercury may be oxidized to Hg2þ, which can then become methylated into methylmercury compounds by chemical or microbiological reactions in the aquatic environment. The intestinal bacterial flora of various animal species, including fish, is also, though to a much lower degree, able to convert ionic mercury into methyl mercuric compounds (Nielsen, 1992). Methylmercury is accumulated by fish and marine mammals and it attains its highest concentrations in large predatory species at the top of the aquatic food chain. By this means, it enters the human diet. Certain microorganisms can demethylate methylmercury, for example, in the gut, while others can reduce Hg2þ to Hg0. Thus, microorganisms are believed to play an important role in the fate of mercury in the environment and in affecting human exposure.

21.4 ENVIRONMENTAL EXPOSURES 21.4.1

Air

In the areas of Europe remote from industrial activity, mean concentrations of total mercury in the atmosphere are reported to be in the range of 2–3 ng/m3 in summer and 3–4 ng/m3 in winter (UNEP, 2002). Mean mercury concentrations in urban air are usually three–fourfold higher. “Hot spots” of mercury concentration exceeding 10,000 ng/m3 have been reported close to industrial emissions or above areas where mercury fungicides have been used extensively. The chemistry of atmospheric mercury is complex and has attracted much research. Except at or near pollution sources, airborne mercury is of limited relevance in regard to human respiratory exposures. The chemical reactions and the partitioning of mercury in gas and aqueous phases, however, are crucial in regard to mercury residence times in the atmosphere and deposition at various latitudes. This is exemplified during the Arctic spring, when the ultraviolet light catalyzes reactions that lead to short-lasting, yet excessive, depositions of mercury (Lindberg et al., 2002). Few data are available on average indoor air pollution due to mercury vapor. Fatalities and severe poisonings have resulted from heating metallic mercury and mercury-containing objects at home. Elemental mercury is sometimes used for certain cultural and religious practices that may involve sprinkling mercury inside, burning it in a candle, or mixing it with perfume; such practices can create exposures that may greatly exceed currently permitted occupational exposure level (Riley et al., 2001). Release of mercury from dental amalgam fillings is otherwise the predominant source of human exposure to inorganic mercury (IPCS, 1991). Energy-saving light bulbs contain mercury, and their increased popularity will no

814

MERCURY

doubt make broken bulbs an important indoor and environmental source of mercury vapor (Appell, 2007). The daily amount of mercury absorbed from the atmosphere into the bloodstream as a result of respiratory exposure in adults is about 32 ng in rural areas and about 160 ng in urban areas. These calculations are based on rural concentrations of 2 ng/m3 and urban concentrations of 10 ng/m3 (absorption rate 80%). Depending upon the number of amalgam fillings, mercury concentrations in inhaled air have ranged up to several thousand ng/m3, and the estimated average daily absorption is thought to vary between 3000 and 17,000 ng (IPCS, 1991). 21.4.2

Diet (Drinking Water and Food)

Mercury in drinking water is usually in the range of 5–100 ng/L, the average value being about 25 ng/L. Although the forms of mercury in drinking water are poorly known, Hg2þ present as complexes and chelates with various ligands is likely the predominant species. The bioaccessibility (i.e., the extent to which a certain mercury complex is available for absorption at the gastrointestinal mucosal surface) may increase or decrease depending on the ligand and the binding strength between the metal and the ligand. Concentrations of mercury in most foodstuffs (EFSA, 2004) are often below the detection limit, and likely to be inconsequential. Freshwater fish, seafood, in general, and marine mammals constitute the dominant sources, mainly in the form of methylmercury compounds (70–90% of the total). The amount of mercury in fish depends on factors such as pH and redox potential of the water, species, age, and size of the fish. The normal concentrations in edible tissues of various species of fish cover a wide range, mostly between 50 and 1400 ng mercury per gram of fresh weight (IPCS, 1990; EFSA, 2004). Large predatory fish, such as pike, swordfish, and tuna, as well as shark, seals, and toothed whales contain the highest average concentrations. Furthermore, exposure might occur from the use of pharmaceuticals, in particular thimerosal, widely applied as a preservative of vaccines and immunogloblins. With up to 100 mg mercury per injection, this preservative caused substantial bolus doses of mercury, especially on a body weight basis, in connection with childhood immunizations. Skin-lightening lotions and soaps used, in Arabian and African countries often contain mercury concentrations of about 1000 mg/kg. Some of these products may even reach concentrations in the percent range. Although mercury may be absorbed through the skin, consumers are usually not warned about the toxic contents. 21.4.3

Relative Significance of Different Routes of Environmental Exposure

Human exposure to the three major forms of mercury present in the environment is summarized in Table 21.1 (based on IPCS, 1991). Although the choice of values given is associated with some uncertainty, the numbers provide a perspective on the relative magnitude of the contributions from various media. Humans may be exposed to additional quantities of mercury occupationally, from living in heavily polluted areas or through the use of skin-lightening creams. The intake from drinking water is about 50 ng mercury per day, mainly as Hg2þ; of which only a small fraction is absorbed. The main pathway of exposure is through the intake of fish and seafood products, mainly in the form of methylmercury. Very high exposures occur in arctic populations, whose diets include marine mammals. Increased levels also occur in Japanese and Mediterranean populations, who frequently eat fish high in the food chain. Exposures are lower in countries, such as the United States, where NHANES III data suggest that 85% of

OCCUPATIONAL EXPOSURES

TABLE 21.1

815

Estimated Average Daily Intake (Retention) of Mercury Compounds Estimated Average Daily Intake (Retention)a in ng of Mercury Per Day Mercury Vapor

Media

b

Air Food Marine Nonmarine Drinking water Dental amalgam Total

Inorganic Mercury Compounds c

Methylmercury

40–200 (30–160)

0

0c

0 0 0

600d (60) 3600 (360) 50 (5)

2400d (2300) — 0

3800–21,000 (3000–17,000) 3900–21,000 (3100–17,000)

0

0

4200 (420)

2400 (2300)

a

Figures in parentheses are the amounts retained that were estimated from the pharmacokinetic parameters (i.e., 80% of inhaled vapor, 95% of ingested methylmercury, and 10% of inorganic mercury are retained). b Assumes an air concentration of 2–10 ng/m3 and a daily respiratory volume of 20 m3. c For the purposes of comparison, it is assumed that in the atmospheric concentrations of species of mercury other than mercury vapor are negligible. d It is assumed that 80% of the total mercury in edible fish tissues is methylmercury and 20% in the form of inorganic mercury compounds. Marine food intake may vary considerably between individuals and across populations.

Americans consume fish at least once a month, 40% once a week, while only 1–2% consume fish or shellfish almost daily. Because fish and seafood is recommended as an essential part of a varied diet, advisories need to identify the types of nutrient-rich fish that are low in mercury to ensure that the benefits exceed the risks. Total dietary mercury intake has usually been measured as part of market basket surveys or as part of specific monitoring. Probabilistic analyses based on dietary questionnaire data and fish analyses suggest that small children, on a body weight basis, may receive a higher exposure than adults (EFSA, 2004). Incomplete information is available on the distribution of high-end intakes from seafood diets, especially among vulnerable population groups, such as pregnant women and children.

21.5 OCCUPATIONAL EXPOSURES Occupational exposure is almost exclusively to inorganic mercury and occurs at chloralkali plants, mercury mines, thermometer factories, fluorescent light tube production plants, refineries, and in dental clinics. High mercury concentrations have been described for all these situations, with considerable variations depending on the working conditions. Some 70,000 workers in the United States were considered exposed to mercury, primarily elemental mercury, but the number is decreasing. Serious mercury exposures may occur in connection with gold mining, especially when gold amalgam is heated. In developing countries, this process is often carried out under field conditions or in small gold vending shops without or with insufficient ventilation. An estimated 10 million workers in Africa, Latin America, and Asia are exposed to high concentrations of elemental mercury through such activities. The impacts on health from these exposures are generally not monitored.

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21.6 KINETICS AND METABOLISM The bioavailability, kinetics, and biotransformation of mercury depend upon its chemical and physical form. 21.6.1

Absorption

21.6.1.1 Elemental Mercury (Hg0) Approximately 80% of inhaled mercury vapor is absorbed via the lungs and retained in the body. Elemental mercury is poorly absorbed in the gastrointestinal tract (less than 0.01% in rats). Increased blood mercury concentrations have been measured in humans, however, after accidental ingestion of several grams of metallic mercury. 21.6.1.2 Inorganic Mercurous (Hgþ) and Mercuric (Hg2þ) Mercury The absorption of inhaled aerosols of inorganic mercury depends on particle size, solubility, and so on (IPCS, 1991). No data have been reported for humans. In dogs, 45% of deposited mercuric oxide aerosols were cleared in less than 24 h, and the remainder cleared with a half-time of 33 days. Ten–fifteen percent of an oral, nontoxic dose of mercuric mercury is absorbed from the gastrointestinal tract in adults and retained in body tissues, but considerable individual variations may exist. In children, the gastrointestinal absorption is likely to be greater. 21.6.1.3 Organic Mercury Human poisoning cases caused by inhalation indicate that a large fraction of these lipophilic compounds are absorbed into the blood. Alkylmercury compounds are absorbed almost completely in the gastrointestinal tract. Certain methylmercury compounds are probably absorbed through the skin. Parenteral uptake of ethylmercury occurs in connection with vaccines preserved with thimerosal. 21.6.2

Distribution

21.6.2.1 Elemental Mercury (Hg0) After exposure to mercury vapor, the element is found in blood as physically dissolved elemental mercury. Within a few minutes, the enzyme catalase oxidizes mercury into mercuric mercury in the erythrocytes. Through this mechanism, the maximum Hg concentration in the erythrocytes can be observed within an hour after a brief exposure to mercury vapor. In contrast, it takes about 10 h for plasma concentrations to peak. Before oxidation, Hg0 readily crosses cell membranes, including the blood/brain barrier and the placental barrier. After oxidation, the Hg2þ ions (or complexes) are distributed in the body via the blood. The kidneys and the brain are the main retention sites for Hg after exposure to mercury vapor, whereas absorbed inorganic mercury salts are mainly deposited in the kidneys. The uptake and/or elimination of mercury after exposure to mercury vapor can be altered by a moderate intake of alcohol, possibly due to inhibition of catalase. In humans, ingestion of alcohol prior to mercury vapor exposure has resulted in significantly reduced blood mercury levels (Hursh et al., 1980). 21.6.2.2 Inorganic Mercurous (Hgþ) and Mercuric (Hg2þ) Mercury The kidneys are the predominant site of inorganic mercury accumulation. After oral exposure, accumulation also occurs in the cells of the mucous membranes of the gastrointestinal tract. A significant

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part of this accumulation, however, never reaches the general circulation as it is eliminated through cell shedding. Mercuric mercury in blood is divided between erythrocytes and plasma in about equal amounts. In erythrocytes, mercury is probably to a large extent bound to sulfhydryl groups of the hemoglobin molecule and possibly also to glutathione. The distribution between different plasma–protein fractions varies with dose and time after exposure. To a limited extent, mercuric mercury crosses the blood–brain and placental barriers. However, mercuric mercury does accumulate in the placenta, fetal membranes, and amniotic fluid. The rate of uptake from blood and different organs varies widely, as does the rate of elimination from different organs. Thus, the distribution of mercury within the body and within organs varies widely with dose and time lapse after absorption. Yet, under all conditions, the dominating mercury pool in the body after exposure to mercuric mercury is the kidney. Inorganic divalent mercury can induce metallothionein, and a large proportion of the mercury in the kidneys is soluble and bound to this molecule. 21.6.2.3 Organic Mercury Methylmercury is distributed via the bloodstream to all tissues in the body. The pattern of tissue distribution is much more uniform than after inorganic mercury exposure. Red cells are an exception; there the concentration is 10–20 times greater than the plasma concentration. Methylmercury readily crosses the blood–brain and placental barriers. In the fetus, methylmercury is accumulated and concentrated, especially in the brain. As with other forms of mercury, the kidneys retain the highest tissue concentration, but the brain still contains about fivefold higher concentrations than blood. Methylmercury accumulates in hair in the process of formation of hair strands, with average concentrations being about 250-fold higher than in the blood. Methylmercury undergoes biotransformation to inorganic mercury by demethylation, particularly in the gut. Ethylmercury is less stable than methylmercury. 21.6.3

Elimination

21.6.3.1 Elemental Mercury (Hg0) After short-term exposure to mercury vapor, about one third of the absorbed mercury will be eliminated in unchanged form through exhalation. The remaining mercury will be eliminated as mercuric mercury mainly through feces. Assuming first-order kinetics for the clearance of urinary mercury after exposure to mercury vapor, the median half-time was found to be 41 days. Blood concentrations can serve as indicators of recent mercury vapor exposure. Speciation, in this case, should be carried out to eliminate possible influence of dietary intake of mercury from contaminated marine food. 21.6.3.2 Inorganic Mercurous (Hgþ) and Mercuric (Hg2þ) Mercury Excretion of absorbed inorganic mercury is mainly via urine and feces, the rates by each pathway being roughly equal. The whole-body half-time in adults is also about 40 days. The elimination of inorganic mercury follows a complicated pattern, with biological half-times differing according to the tissue and the time after exposure. Thus, mercury concentrations in critical organs may remain high even after having dwindled in urine and blood. Hence, at present, there are no general and suitable indicator media that will reflect concentrations of inorganic mercury in the critical organs, the brain or kidney, under different exposure conditions (IPCS, 1991).

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21.6.3.3 Organic Mercury Mercury excretion after methylmercury exposure is predominantly via the feces. Methylmercury is slowly demethylated in the gut, and the enterohepatic recirculation of methylmercury explains that most, if not all, of the mercury excreted is in the demethylated inorganic form. Some elimination also occurs via urine. The whole-body half-time of methylmercury is generally about 45 days, although higher estimates have also been published. Laboratory animal studies have shown that following acute dosage with methylmercury, blood mercury concentrations will initially reflect organ concentrations reasonably well. Henceforth, an increasing fraction of the body burden will be in the brain, muscles, and kidney. The blood concentration might be a useful indicator of the body burden of mercury while the erythrocyte mercury concentration is more specific for methylmercury exposure. Accordingly, if exposure to mercury vapor or other inorganic mercury compounds is suspected, mercury should be speciated or a serum sample analyzed. Mercury in hair, when measured along the length of a hair strand, has also been used as an indicator of past blood levels. When using hair as an indicator, it is important to note the history and condition of the hair. Permanent waving may leach mercury from the hair while exogenous mercury can increase its concentration. Cord blood (or cord tissue) measurement is the best method to establish prenatal exposure levels.

21.7 HEALTH EFFECTS 21.7.1

Acute and Local Effects

Acute poisoning with mercury vapor may cause a severe airway irritation, chemical pneumonitis, and, in severe cases, pulmonary edema. Ingestion of inorganic compounds may cause gastrointestinal corrosion and irritation, such as vomiting, bloody diarrhea, and stomach pains. Subsequently, shock and acute kidney dysfunction with uremia may ensue. Local irritation may develop following cutaneous exposure to mercury compounds, which are among the most common allergens in patients with contact dermatitis. 21.7.2

Chronic and Systemic Effects

Chronic intoxication may develop as early as a few weeks after the onset of a mercury exposure. More commonly, however, the exposure has lasted for several months or years, yet early diagnosis is thwarted by the lack of recognition of subtle effects. The symptoms depend on the degree of exposure and the kind of mercury in question. They mainly involve the oral cavity, the peripheral and central nervous system, and the kidneys. As the elemental mercury present in vapor is oxidized to mercuric mercury in the blood, the non-neurotoxic effects of absorbed mercury vapor and other inorganic mercury compounds will be similar. 21.7.2.1 Elemental Mercury (Hg0) Severe exposure to inorganic mercury causes an inflammation of gingiva and oral mucosa, which become tender and bleed easily. Salivation is increased, most obviously so in subacute cases. Often the patient complains of a metallic taste in the mouth. Especially when oral hygiene is bad, a gray border is formed on the gingival edges. In exposures to mercury vapor, the central nervous system is the critical organ, and the classic triad of symptoms includes erethism, intention tremor, and the gingivitis described

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above. The fine intention tremor of fingers, eyelids, lips, and tongue may progress to spasms of arms and legs. A jerky micrographia is typical as well. The changes in the central nervous system result in psychological effects known as erethism: restlessness, irritability, insomnia, concentration difficulties, decreased memory and depression, sometimes in combination with shyness, unusual psychological vulnerability, and anxiety. “Micromercurialism” is a term used to denote an early stage of erethism in which decreased memory, dizziness, and irritability are most prominent. While recent data do not support an association between dental fillings and deleterious effects some patients believe that their symptoms are linked to amalgam restorations (Bellinger et al., 2006). Induction of minimal tremor by mercury vapor has been reported at urinary excretion levels of 50 mg/L (0.25 mmol/L) and above. Data concerning the effects of mercury vapor on early stages of the human life cycle are limited. While some information is available with regard to effects on pregnancy and birth in women occupationally exposed to mercury vapor, a dose–response relationship has not been established. In children, “pink disease” may occur, as described below. 21.7.2.2 Inorganic Mercurous (Hgþ) and Mercuric (Hg2þ) Mercury The target organ following long-term exposure causing no acute toxicity are the kidneys. In general, the early renal effects of mercury appear to be reversible after cessation of exposure. Nephrotoxic effects include proximal tubular damage, as indicated by an increased excretion of small proteins in the urine (e.g., beta2-micro globulin). Experimental studies suggest that glomerular damage is caused by an autoimmune reaction to mercury complexes in the basal membrane. This mechanism of action, however, has not been confirmed for humans. In children, a different syndrome is seen, the so-called pink disease or acrodynia, diagnosed most frequently in children treated with teething powders, which contained calomel. It has also been occasionally seen in children who had inhaled mercury vapor (e.g., from broken thermometers) (Agocs et al., 1990). A generalized eruption develops and the hands and feet show a characteristic, scaly, reddish appearance. In addition, the children are irritable, sleep badly, fail to thrive, sweat profusely, and have photophobia. This condition was extremely common until 30 years ago, when the etiology was finally found and teething powders were phased out. 21.7.2.3 Organic Mercury Intoxications with alkoxialkyl or aryl compounds are similar to intoxications with inorganic mercury compounds due to their relatively unstable state. Alkyl mercury compounds, such as methylmercury, result in a different syndrome. The earliest symptoms in adults are paresthesias in the fingers, the tongue, and the face, particularly around the mouth. Later on, disturbances occur in the motor functions, resulting in ataxia and dysphasia. The visual field is decreased and, in severe cases, the result may be total blindness. Similarly, impaired hearing may progress to complete deafness. This syndrome appeared as Minamata disease in Japan as a result of methylmercury contamination from a local factory. Epidemics also occurred when methylmercurytreated seed grain was used for baking or animal feed in Iraq and elsewhere. Children and the fetus are more susceptible to the toxic effects of methylmercury than are adults, and congenital methylmercury poisoning may result in a cerebral palsy syndrome, even though the mother remains healthy or suffers only minor symptoms due to the exposure. In populations with a high consumption of fish or marine mammals, methylmercury intakes may approach the levels that resulted in such serious disease in Japan and Iraq. Recent evidence from long-term follow-up of a Faroese birth cohort has shown that prenatal

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exposure to methylmercury may result in neuropsychological and neurophysiological deficits that are detectable through adolescence (Grandjean et al., 1997; Murata et al., 2004; Debes et al., 2006; Budtz-Jørgensen et al., 2007). In adults, the earliest effects, such as paresthesias, appear to occur when blood concentrations are above 200 mg/L (1 mmol/L). Recent epidemiological studies suggest that adverse cardiovascular effects may occur at much lower exposures than are prevalent among people regularly eating seafood (Virtanen et al., 2005). Although the implications of these findings are not yet clear, they may suggest that methylmercury can be a toxic risk to the population at large, and that benefits of eating seafood must take into account that mercury exposures need to be minimized. Developmental delays appear to be related to maternal hair mercury concentrations of 1–3 mg/g (Grandjean et al., 1997) (i.e., cord blood concentrations of 4–12 mg/L). Despite several years of research, the evidence on the possible adverse health effects of thimerosal in vaccines remains unclear (Institute of Medicine, 2004; Geier and Geier, 2006). Sufficient evidence supports that methylmercury chloride is carcinogenic to experimental animals. In the absence of comprehensive epidemiological data, methylmercury is considered a possible human carcinogen (class 2B) (IARC, 1994). The U.S. EPA has classified both inorganic mercury compounds and methylmercury as possible human carcinogens.

21.8 PREVENTION Prevention should start at the source. For example, the European Union has recently enacted a ban on mercury exports and a variety of mercury uses are being phased out, such as mercury in household thermometers. Of important nonindustrial sources, batteries and fluorescent light bulbs are recycled in many countries. Mercury exposures from dental amalgam fillings should be minimized, and suitable alternative restorative materials are now available for most purposes. Pollution abatement should also focus on point sources, such as coal-fired power plants and incinerators. When mercury emissions from such sources were controlled in Florida and Massachusetts, methylmercury contamination of local freshwater fish significantly decreased after a few years. However, on a global scale, much still remains to be done. This is particularly the case with regard to pollution from burning of mercurycontaining coals in East Asia. In regard to occupational exposures, WHO has recommended that long-term mercury vapor exposures should be limited to a time-weighted average (TWA) limit of 25 mg/m3, a value that has also been adopted by the ACGIH as a threshold limit value. The corresponding TWA for inorganic mercury is 50 mg/m3. Biological monitoring is crucial in the diagnosis of mercury exposure and in the control of occupational exposure levels. Although blood concentrations are highly useful, they do not reflect mercury retained in the brain, where mercury from vapor inhalation has a half-life of several years. Urine levels are usually preferred as an indicator of occupational exposures to inorganic mercury species, and a limit of 50 mg Hg/g creatinine (28 mmol/mol creatinine) has been recommended (IPCS, 1991). For consumers, recent exposure can be ascertained by analysis of hair or blood. In the United States, increased mercury concentrations in freshwater fish have prompted fish advisories in almost every single state. However, current advisories may not be sufficient to help consumers obtaining the optimal nutritional benefits from fish and seafood, while minimizing methylmercury exposure. Current limits of 0.5 mg/g for fish in general and

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1.0 mg/g for large fish, such as swordfish and tuna, are not based on a detailed risk assessment, but reflect methylmercury concentrations prevalent in seafood. A healthy diet that includes two fish dinners per week must be based on seafood that averages a mercury concentration of no more than 0.1 mg/g to avoid exceeding the reference dose (RfD) of 0.1 mg/kg body weight per day defined by the U.S. Environmental Protection Agency (2001). Popular and nutritious fish with low mercury concentrations include salmon, sardine, and flounder. Exposure limits for methylmercury have been revised downward to protect sensitive life stages (National Research Council, 2002; JECFA, 2003). The major limits are the reference dose (RfD) of 0.1 mg/kg body weight per day (National Research Council, 2002) and the provisional tolerable weekly intake (PTWI) of 1.6 mg/kg body weight per week (JECFA, 2003) correspond to hair mercury concentrations of approximately 1–2 mg/g. Data on blood mercury concentrations in the U.S. general population suggest that between 5 and 10% of different population sections exceed the RfD.

REFERENCES Agocs MM, Etzel RA, Parrish RG, Paschal DC, Campagna PR, Cohen DS, Kilbourne EM, Hesse JL (1990) Mercury exposure from interior paint. N. Engl. J. Med. 323:1096–1101. Appell D (2007) Toxic bulbs. Sci. Am. 297 (4):30–31. ATSDR (Agency for Toxic Substances and Disease Registry)(1994) Toxicological Profile for Mercury (update), TP-93/10, Atlanta, GA. Available at http://www.atsdr.cdc.gov/toxprofiles/tp46.html (accessed March 12, 2006). Bellinger DC, Trachtenberg F, Barregard L, Tavares M, Cernichiari E, Daniel D, McKinlay S (2006) Neuropsychological and renal effects of dental amalgam in children: a randomized clinical trial. JAMA 295:1775–1783. Budtz-Jørgensen E, Grandjean P, Weihe P (2007) Separation of risks and benefits of seafood intake. Environ. Health Perspect. 115:323–327. Debes F, Budtz-Jørgensen E, Weihe P, White RF, Grandjean P (2006) Impact of prenatal methylmercury toxicity on neurobehavioral function at age 14 years. Neurotoxicol. Teratol. 28:363–375. EFSA (European Food Safety Authority) (2004) Opinion of the Scientific Panel on Contaminants in the Food Chain on a Request from the Commission Related to Mercury and Methylmercury in Food (EFSA-Q-2003-030), Brussels. Available at http://www.efsa.eu.int/science/contam/contam_opinions/259_en.html (accessed March 12, 2006). Geier DA, Geier MR (2006) An evaluation of the effects of thimerosal on neurodevelopmental disorders reported following DTP and Hib vaccines in comparison to DTPH vaccine in the United States. J. Toxicol. Environ. Health A 69:1481–1495. Grandjean P, Weihe P, White RF, Debes F, Araki S, Yokoyama K, Murata K, Sørensen N, Dahl R, Jørgensen PJ (1997) Cognitive deficit in 7-year-old children with prenatal exposure to methylmercury. Neurotox. Teratol. 19:417–428. Harris RC, Rudd JW, Amyot M, Babiarz CL, Beaty KG, Blanchfield PJ, Bodaly RA, Branfireun BA, Gilmour CC, Graydon JA, Heyes A, Hintelmann H, Hurley JP, Kelly CA, Krabbenhoft DP, Lindberg SE, Mason RP, Paterson MJ, Podemski CL, Robinson A, Sandilands KA, Southworth GR, St Louis VL, Tate MG (2007) Whole-ecosystem study shows rapid fish-mercury response to changes in mercury deposition. Proc Natl Acad Sci USA 104:16586–16591. Hursh JB, Greenwood MR, Clarkson TW, Allen J, Demuth S (1980) The effect of ethanol on the fate of mercury vapor inhaled by man. J. Pharmacol. Exper. Therap. 214:520–527.

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IARC (1994) Monographs on the Evaluation of Carcinogenic Risk to Humans, Vol. 58. Mercury and Mercury Compounds, Lyon. Institute of Medicine (2004) Immunization Safety Review: Vaccines and Autism. Washington: National Academy Press. IPCS (International Programme on Chemical Safety) (1990) WHO Task Group on Environmental Health Criteria for Methylmercury (EHC 101), Geneva: WHO. IPCS (International Programme on Chemical Safety) (1991) WHO Task Group on Environmental Health Criteria for Inorganic Mercury (EHC 118), Geneva: WHO. JECFA (Joint FAO/WHO Expert Committee on Food Additives) (2003) Sixty-first meeting, Rome, 10–19 June 2003. Summary and Conclusions. Available at ftp://ftp.fao.org/es/esn/jecfa/jecfa61sc. pdf (accessed March 19, 2006). Lindberg SE, Brooks S, Lin CJ, Scott KJ, Landis MS, Stevens RK, Goodsite M, Richter A (2002) Dynamic oxidation of gaseous mercury in the Arctic troposphere at polar sunrise. Environ. Sci. Technol. 36:1245–1256. Murata K, Weihe P, Budtz-Jørgensen E, Jørgensen PJ, Grandjean P (2004) Delayed brainstem auditory evoked potential latencies in 14-year-old children exposed to methylmercury. J. Pediatr. 144:177–183. National Research Council (2002) Toxicological Effects of Methylmercury. Washington: National Academy Press. Nielsen JB (1992) Toxicokinetics of mercuric chloride and methylmercuric chloride in mice. J. Toxicol. Environ. Health 37:85–122. Rice G, Hammitt JK (2005) Economic valuation of human health benefits of controlling mercury emissions from U.S. coal-fired power plants, NESCAUM. Available at http://www.nescaum.org/ topics/mercury-control-technology (accessed March 12, 2006). Riley DM, Newby CA, Leal-Almeraz TO, Thomas VM (2001) Assessing elemental mercury vapor exposure from cultural and religious practices. Environ. Health Perspect. 109:779–784. UNEP (United Nations Environment Programme) (2002) Global Mercury Assessment Report, Geneva. Available at http://www.chem.unep.ch/mercury/Report/Final%20Assessment%20report.htm (accessed March 12, 2006). U.S. EPA (Environmental Protection Agency) (2001) Water Quality Criterion for the Protection of Human Health: Methylmercury. Publication EPA-823-R-01-001, Washington, D.C. Available at http://www.epa.gov/waterscience/criteria/methylmercury/document.html (accessed March 12, 2006). Virtanen JK, Voutilainen S, Rissanen TH, Mursu J, Tuomainen TP, Korhonen MJ, Valkonen VP, Seppanen K, Laukkanen JA, Salonen JT (2005) Mercury, fish oils, and risk of acute coronary events and cardiovascular disease, coronary heart disease, and all-cause mortality in men in eastern Finland. Arterioscler. Thromb. Vasc. Biol. 25:228–233.

22 NITROGEN OXIDES Richard B. Schlesinger

22.1 INTRODUCTION Oxides of nitrogen (NOx), so called because they consist of various chemical species, many of which are interconvertible, can exist in the atmosphere as either gases/vapors or particles. The former includes nitric oxide (NO), nitrogen dioxide (NO2), nitrous oxide (N2O), and occasionally, nitrogen trioxide (NO3), dinitrogen trioxide (N2O3), dinitrogen tetroxide (N2O4), and dinitrogen pentoxide (N2O5), while the latter includes nitrate (NO3 ) salts. Species that may exist in either a particulate or gaseous state are nitric acid (HNO3) and nitrous acid (HONO). Nitric oxide and nitrogen dioxide are the most important of the NOx in terms of public health concern since they are often present in the atmosphere in significant concentrations and are quite chemically reactive. Although N2O is also ubiquitous, being released due to natural biological processes in soil, it is not involved in chemical reactions in polluted air. Most other NOx, if found at all, are present at very low concentrations. Two possible exceptions from a health standpoint are HNO3 and inorganic nitrates.

22.2 SOURCES Ambient atmospheric NOx derive from both natural sources, such as forest fires, organic decay, and lightning, and from anthropogenic activities that involve high temperature combustion processes in both mobile and stationary sources. The major mobile source is motor vehicles, while the major stationary source is electric power generation using fossil fuels, with industrial combustion processes being a close second. From a global perspective, however, the total mass of emissions released from natural sources is much greater than that from human activities. NOx, primarily NO2, can also be an important indoor pollutant. The


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main indoor source is the use of unvented or improperly vented natural gas or other fossil fuel-fired appliances, such as stoves and heaters. During combustion processes, nitrogen derived from the combustion air and/or the fuel being consumed reacts with atmospheric oxygen. Although most of the resulting NOx produced is initially in the form of NO, this is generally rapidly oxidized to NO2, with the conversion rate depending upon a number of factors, including the concentration of NO, temperature of the combustion process, and distance from the emission zone. Several reaction pathways are possible. While simple oxidation involving molecular oxygen (O2) is the primary one for NO2 production in combustion gas effluents, it does not play a major role in the ambient atmosphere since transformations via other pathways occur at faster rates. Thus, in air containing other reactive chemical species, for example, ozone, irradiation by sunlight can catalyze photochemical reactions leading to the very rapid formation of NO2. HNO3 is also a product of the photooxidation cycle of polluted air but, along with HONO, can additionally derive from primary emissions released by mobile sources. The major production pathway involves reaction between the hydroxyl radical (OH), formed within the smog cycle, with NO2. Other routes, which are potentially important at night, involve reactions between N2O5 with water or nitrate radicals with volatile organics, or production in droplets containing both hydrogen ion (Hþ) and nitrate (NO3 ). Because of its high saturation vapor pressure, HNO3 generally exists as a vapor under most ambient conditions, for example, within photochemical smog, where levels generally peak during daytime hours (Ellestad and Knapp, 1988). Within acidic fogs, however, HNO3 may be found in the particulate state (Jacob et al., 1985). Similarly, HONO can be found in ambient air both as a primary product from combustion sources and as a secondary product of photochemical smog reactions. HNO3 may also be produced indoors via reaction of ozone with NO2, water vapor, and volatile organics (Weschler et al., 1992). Water on indoor surfaces can react with NO2 to form HONO, which can then be released into indoor air as gas phase acid (Dubowski et al., 2004). Nitrate salts may be formed in the atmosphere via various pathways, many of which involve gaseous HNO3. For example, ammonium nitrate results from the homogeneous reaction between nitric acid and atmospheric ammonia. Nitrates may also be formed by heterogeneous reactions involving NO2 or NO and water droplets, or HNO3 vapor and dust or sea salt particles.

22.3 NITROGEN DIOXIDE 22.3.1

Exposure

22.3.1.1 Atmospheric Concentration Outdoor levels of NO2 are often directly related to motor vehicle emissions and traffic density around busy roadways and, along with particulate matter (PM) and various organics, NO2 is considered to be a good indicator of the complex particulate–gas mixture that derives from vehicular traffic (Gauderman et al., 2000, 2002; McConnell et al., 2003; Seaton and Dennekamp, 2003). Outdoor concentrations in urban areas are generally characterized by two daily peaks related to traffic patterns in the morning and afternoon. In areas having significant stationary sources, the pattern is characterized by a baseline NO2 level superimposed upon which are higher spikes occurring on an irregular basis. In those areas not impacted by significant local sources,

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NO2 has little variation on an hourly basis throughout the day, unless there is transport of NO2 into the region. The average daily l-h maximum outdoor concentration of NO2 in urban regions across the United States was 0.03 ppm; levels in nonmetropolitan or rural areas would tend to be lower. The 24-h average is about half this value (U.S. EPA, 2008). Data from various rural, suburban, and urban sites in Southern California indicated 8-year (1994–2001) mean concentrations ranging from 0.003 to 0.038 ppm, with annual means at the most rural sites ranging from 0.003 to 0.02 ppm. While, as noted, natural emissions far outweigh those from anthropogenic sources on a total mass basis, the former are distributed over a wider area; this results in very low background levels due to natural sources. Nitrogen dioxide is also an indoor air pollutant, deriving from combustion sources such as gas-fired ranges, kerosene heaters, and improperly or unvented gas space heaters. Nitrogen oxides are also major components of smoke derived from the burning of tobacco products. Cigarette smoke contains high levels of NO, which is oxidized to NO2 as the smoke ages. The indoor/outdoor concentration ratio for NO2 in the absence of significant indoor sources is 0.5–0.6, but is often >1 when such sources are present (Berglund, 1993). Indoor levels vary widely depending upon the strength of the specific sources and the degree of ventilation. Furthermore, because combustion from indoor sources tends to be episodic, fairly high short-term peaks are possible. Daily (24-h average) levels of NO2 in homes using gas-fired ranges or heaters can range between 0.05 and 0.5 ppm, but short-term peaks can exceed 1 ppm (U.S. EPA, 1993; Spengler and Cohen, 1985; Goldstein et al., 1988). 22.3.1.2 Exposure Assessment From a health standpoint, the only relevant route of exposure to NOx is via inhalation and, from the above discussion, it is evident that such exposures can occur in numerous settings, which include residential areas, transportation vehicles, as well as the outdoor atmosphere. The integrated exposure is, therefore, the sum of the individual exposures over all possible time intervals and for all of these different environments. Such exposure can be assessed either by direct methods, which include biomarkers and personal monitoring, and indirect methods, which involve measurement of pollutant levels at monitoring sites and the use of mathematical models for the estimation of actual individual or population exposures. There is currently no accepted biomarker for exposure to NO2. Some suggested ones have included urinary hydroxyproline excretion (Adgate et al., 1992), the NO-heme protein complex in bronchial lavage (Maples et al., 1991) and 3-nitrotyrosine in urine (Oshima et al., 1990). However, because of their lack of sensitivity and/or specificity, these have not been shown to be practical for assessing environmental NO2 exposures. Outdoor measures of NO2 levels, while related to and contributing to total exposure, are poor predictors of total personal exposures for most people. Because indoor concentrations are often greater than those outdoors, indoor exposure is commonly the main contributor to total exposure, and actual personal exposures to NO2 may differ from what would be assumed based upon ambient outdoor air measures (Linaker et al., 1996). Thus, indoor residential levels are generally a much better predictor of personal exposure, explaining over 50% of the variation in such exposure. However, it should be borne in mind that this is a generalization, and there are likely to be selected groups of people for which indoor levels are not a good predictor of total exposure due to greater percentages of time spent in other significant NOx-containing environments, especially those occupationally exposed or those living near heavily traveled roadways.

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TABLE 22.1

Exposure Limits for Nitrogen Oxides

Nitric oxide TLVa PELb RELc IDLHd

25 ppm 25 ppm 25 ppm 100 ppm

Nitrogen dioxide NAAQSe PELf RELg EEGLh TLVa STELi IDLHd

0.053 ppm 5 ppm 1 ppm 1 ppm 3 ppm 5 ppm 20 ppm

a

Threshold limit value (ACGIH; time weighted average for an 8-h work day and a 40-h work week). Permissible exposure limit (OSHA; time weighted average for an 8-h work day). c Recommended exposure limit (NIOSH; time weighted average for an 8-h work day). d Immediately dangerous to life and health (NIOSH; 30-min average). e National ambient air quality standard (USEPA; annual average). f Permissible exposure limit for general industry (OSHA; ceiling for 15 min). g Recommended exposure limit (NIOSH; ceiling for 15-min exposure). h Emergency exposure guidance level (NAS; 1-h exposure). i Short-term exposure limit (ACGIH; ceiling for 15-min exposure). b

22.3.1.3 Exposure Limits There are various ambient and occupational exposure limits for NOx. Some of the major ones are listed in Table 22.1. 22.3.2

Dosimetry

Up to 90% of the NO2 inspired during normal respiration can be removed within the human respiratory tract (Wagner, 1970). Estimates of regional uptake for the upper respiratory tract (i.e., airways proximal to the trachea) based upon laboratory animal studies range from 28 to 90% of the amount inhaled (Cavanagh and Morris, 1987; Dalhamn and Sjoholm, 1963; Yokoyama, 1968; Vaughan et al., 1969; Kleinman and Mautz, 1989), while that for the lungs range from 36 to 90% of the amount entering the trachea (Postlethwait and Mustafa, 1981; Kleinman and Mautz, 1989). Specific ventilatory factors influence the extent of uptake. Thus, more NO2 will be absorbed in the upper respiratory tract during nasal breathing than during oral breathing (Kleinman and Mautz, 1989), implying that the latter would allow a greater percentage of inhaled NO2 to reach the lungs. Increased ventilation, such as due to exercise, may alter regional gas distribution from that occurring at rest by reducing NO2 uptake in the upper respiratory tract and tracheobronchial tree and, thus, increasing the amount of NO2 delivered to and absorbed in the respiratory (alveolar) region of the lungs (Miller et al., 1982; Overton, 1984; Kleinman and Mautz, 1989; Wagner, 1970; Mohsenin, 1994). Within the lungs, inhaled NO2 can be absorbed throughout the entire tracheobronchial tree and respiratory region, although the major dose to tissue is delivered at the junction between the conducting and respiratory airways (Miller et al., 1982; Overton, 1984). Regardless of the site of initial contact with airway surfaces, a primary determinant of

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NO2 uptake is surface reactivity, that is, direct interaction with airway lining fluid and/or cellular components (Postlethwait and Bidani, 1990). Potential substrates include oxidizable chemical species, such as amino acids, proteins, and unsaturated fatty acids (Hood et al., 1993), resulting in the production of nitrite ion or various radicals (Postlethwait and Mustafa, 1981; Saul and Archer, 1983; Postlethwait and Bidani, 1989, 1990), which can then interact with the epithelium or rapidly pass into the bloodstream and undergo other chemical reactions in extrapulmonary sites, for example, oxidation to nitrate by interaction with hemoglobin in red blood cells (Parks et al., 1981; Oda et al., 1981; Kosaka et al., 1979; Case et al., 1979). Nitric and nitrous acids, as well as their nitrate salts, have been detected in blood and urine following exposure to NO2 (U.S. EPA, 1993; Garn et al., 2003). Antioxidants present within airway lining fluid can react with deposited NO2, potentially modulating its toxicological impact (Kelly et al., 1996). Any NO2 that dissolves in airway fluids could result in the production of nitric and nitrous acids, with any subsequent toxicity likely due to the hydrogen or nitrite ions (Goldstein et al., 1977, 1980). It is, however, likely that both oxidative and nonoxidative mechanisms are involved in toxicity from inhaled NO2. 22.3.3

Health Effects––Epidemiology

Epidemiological studies have attempted to assess the potential role of exposure to NO2 in producing adverse human health effects. Many studies relate health end points to outdoor concentrations, but the current trend is to provide better measures of actual personal exposures, which, as noted, can be reflections of strong indoor sources. Some studies have used NO2 as the only pollutant, while others have used NO2 as a general marker for pollution derived from motor vehicles. A major problem, however, is, as noted above, a close association between NO2 and other pollutants, especially PM, derived from the same combustion sources, making it often difficult to determine any independent effects due solely to NO2. While the robustness of some epidemiological studies are affected by a lack of reliable estimates of actual NOx exposure conditions, inadequate sample size, inadequate compensation for the effects of covariates, and/or misclassification of health end points, they do provide a linkage between controlled exposure (toxicology) studies and “real-world” exposure of humans. These studies have examined the relationship between acute exposure and effects, as well as responses to long-term exposure. Ambient NO2 has been related to increased mortality in some evaluations (e.g., Wietlisbach et al., 1996; Sunyer et al., 1996; Anderson et al., 1996). A meta-analysis examining daily mortality that incorporated studies published between 1982 and 2000 and that used data from 1958 to 1999 (Stieb et al., 2002, 2003) indicated an overall effect estimate for overall mortality from NO2 alone to be 2.8% (with a 95% confidence interval (CI) of 2.1– 3.5%) per 0.024 ppm NO2 (24-h mean); this value was reduced to 0.9% (CI ¼ 0.1–2.0) in a multipollutant model that included PM as well, highlighting the difficulty in evaluating effects due to NO2 alone. A multicity study found a 1.3% increase in daily mortality (95% CI ¼ 0.9–1.8) per 50 mg/m3 (0.028 ppm) 1-h maximum concentration increase in NO2 (Touloumi et al., 1997). A study examining European cities (Katsouyanni et al., 2001) noted a higher daily mortality in those having higher NO2. Analysis of daily air pollution and mortality in 90 cities in the United States (Dominici et al., 2003) showed a significant increase in daily mortality ranging from 0.3 to 0.4% per 0.010 ppm (using a 1-day lag between concentration and response), but this disappeared when adjusted for PM and ozone.

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Most studies of acute responses to NO2 used indices of respiratory illness and/or changes in pulmonary mechanical function to assess the health consequences of exposure. Some examples of representative results obtained in such studies are provided. In a classic series of surveys conducted in a number of cities in the United States selected to represent a range of outdoor air quality (Harvard Six-Cities Air Pollution Health Study), grade school age children within each community were followed for several years by reporting on questionnaires and by annual measurements of pulmonary function. Outdoor NO2 levels were measured at various sites within each community and indoor levels were also measured in selected households. While results of this study from 1974 to 1977 on over 8000 children aged 6–10 years indicated a significant increase in the rate of respiratory illness before age 2 in homes with gas-fired stoves compared to those with electric stoves (Speizer et al., 1980), a later examination of the same communities over a longer time period did not show any NO2related increase in respiratory illness (Ware et al., 1984). A further analysis of over 5000 children aged 7–11 years during the period 1983–1986 noted marginal significance for physician-diagnosed respiratory illness prior to age 2 in homes using gas-fired stoves compared to those using electric stoves (Dockery et al., 1989). When pulmonary mechanical indices were evaluated in the above-mentioned children (Ware et al., 1984), gas stove use was associated with significant reductions in parameters of expiratory flow (FEV1, FVC) in a first examination, but no such relationship was found in a subsequent evaluation. Certain subpopulations, based upon age or preexisting disease state, or both, may be more susceptible to effects of NO2 than are others. A borderline significant effect was noted between peak flow reduction in healthy children residing in homes having gas stoves, while a much stronger association was noted in asthmatics (Lebowitz et al., 1985). Children with asthmatic symptoms appeared to be more susceptible to reduced lung function when outdoor average NO2 concentrations exceeded a certain level (0.02 ppm), but no such effect was found with children having no asthmatic symptoms (Moseler et al., 1994). A relationship between outdoor levels of NO2 and common respiratory symptoms (e.g., cough, sore throat, etc.) in children up to 5 years of age was noted in one study (Gnehm et al., 1988), while another found an association between NO2 exposure and wheeze in females, but not in males, aged 4 months to 4 years (Pershagen et al., 1995). Braun-Fahrlander et al. (1992) examined symptoms in children in relation to outdoor and indoor levels of NO2. While the incidence of symptoms was not associated with either indoor or outdoor levels, the duration of increased symptoms was associated with outdoor NO2 concentration. In studies of the effect of NO2 on respiratory health in 6–9-year-old children, personal exposures to NO2 were measured, as were indoor levels in the home (Houthuijs et al., 1987; Brunekreef et al., 1987). The prevalence of lung disease was found to be associated with the presence of unvented gas water heaters, with weekly average exposures estimated at 0.021 ppm. On the contrary, Dijkstra et al. (1990) found no association between respiratory symptoms with indoor NO2 measurements in homes. Koo et al. (1990) used personal samplers to monitor NO2 exposure in children aged 7–13 years in Hong Kong. No association was noted between exposure levels (means ranged from 0.013 to 0.023 ppm for a 1-week period) and respiratory symptoms, such as wheeze, running nose, or cough. The effects of both indoor and outdoor air pollution on respiratory illness in a cohort of primary school children indicated a gradient of increased respiratory symptoms with increasing indoor levels of NO2 in homes with gas stoves (Melia et al., 1977). A later assessment also indicated some increase in relative risk in homes with gas stoves, but this was not a consistent finding (Melia et al., 1979). In this case, levels of NO2 measured in bedrooms of homes having gas stoves ranged from 0.003 to 0.017 ppm. In another study of children

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aged 5–6 years, no significant relationship was noted between levels of NO2 and the prevalence of respiratory illness (Melia et al., 1982); levels of NO2 in the bedrooms of homes with gas stoves were 0.005–0.029 ppm. Other attempts to relate gas stove use in homes to acute respiratory illness, respiratory symptoms or indices of reduced lung function for various age populations have had mixed results, with some studies reporting no association and others reporting some relationship (Keller et al., 1979; Comstock et al., 1981; Schenker et al., 1983; Dodge, 1982; Ekwo et al., 1983). A strong association was found between NO2 and asthma admissions to hospital for children under 14 years of age in three European cities (Sunyer et al., 1997), while significant effects of NO2 on doctors visits or hospital admissions for asthma were found in children in various areas of the world (Medina et al., 1997; Anderson et al., 1998; Morgan et al., 1998; Lee et al., 2002; Lin et al., 2002). A stronger association was found to exist between NO2 and asthma symptoms in children in London than for adults in the same area (Hajat et al., 1999). In a study examining respiratory symptoms in adult women and children aged 13 years and younger (Berwick et al., 1984), indoor NO2 levels were measured in homes. Children under the age of 7 years exposed to 0.016 ppm were found to be at an increased risk of upper and lower respiratory tract symptoms compared to those who were not so exposed. No increased risk was found in older children or adults. In a later study (Samet et al., 1993), no association was found between indoor levels of NO2 and either the incidence or duration of respiratory illness in infants examined during their first 18 months of life. A meta-analysis (Hasselblad et al., 1992) of studies using indoor NO2 levels suggested a relationship between incidence of lower respiratory tract symptoms and chronic exposure in children less than 12 years of age, while no effect on lower respiratory tract illness during the first year of life was seen in relation to indoor NO2 in another study (Sunyer et al., 2004). A meta-analysis using data from Australia and New Zealand (Barnett et al., 2005) noted an association between all respiratory-related hospital admissions, as well as admissions specifically for asthma, pneumonia, and acute bronchitis in children grouped into various age ranges. While both PM and NO2 were noted to be associated with total respiratory admissions in children in the 1–4-year age group, the largest effect was noted for NO2. In older children, aged 5–14 years, an association was also found with PM and NO2, but with the latter showing a larger effect. However, the greatest association was noted for asthma admissions increase, related to an increase of about 5 ppm in the 24-h mean concentration of NO2. However, when PM was controlled for, the effect of NO2 in the younger children was attenuated, but that in the older age group was not. Finally, a number of studies have noted associations between NO2 and various symptoms, such as cough, wheeze, and shortness of breath, in children with asthma (e.g., Just et al., 2002; Segala et al., 1998; Quackenboss et al., 1991; Mortimer et al., 2002). In an examination of the relationship between air pollutants and emergency room visits in Atlanta, GA (Peel et al., 2005), the effect estimate for respiratory admissions due to a 1-h exposure to 0.020 ppm was found to be 1.6% (95% CI ¼ 0.6–2.7) for all respiratory admissions, and 1.9% (95%CI ¼ 0.6–3.1) for upper respiratory infections only. While the above-mentioned effects were attenuated in multipollutant models that also considered PM and CO, the effect of NO2 on emergency room visits specifically for asthma was not, and NO2 showed the strongest association with these visits of all pollutants in the model. Asthma may not be the only disease that can predispose to enhanced effects from exposure. Some studies (e.g., D’Ippoliti et al., 2003; Burnett et al., 1997; Wong et al., 1999) noted an association between NO2 and admissions for cardiovascular disease but, as above, effect estimates were often modulated when other pollutants, especially PM, were

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considered in the model as well. Some studies strongly suggest (Mann et al., 2002; Metzger et al., 2004) that individuals with ischemic heart disease and accompanying congestive heart failure and/or arrythmia may be a group especially sensitive to effects of motor vehicle derived pollution, which includes NO2. In an examination of air pollution and emergency room visits for cardiovascular causes in Atlanta, GA (Metzger et al., 2004), there was found to be a significant effect of NO2 on cardiovascular emergency room visits that remained so even after adjusting for PM. In an examination of air pollutants and congestive heart failure (Wellenius et al., 2005), effects with NO2 were modulated when considering CO and PM, but there was a conclusion that the general mix of pollution from motor vehicles was responsible for the observed effects. Finally, daily outdoor concentrations of NO2 were associated with emergency room admissions due to cerebrovascular disease and short-term ischemic attacks (Ponka and Virtanen, 1996). The discussion above involved acute responses to exposure. A few studies have examined the prevalence and/or incidence of asthma or allergic airway disease related to long-term exposure to NO2. These often show conflicting results, and there is no unequivocal evidence that chronic exposure to NO2 will increase these health outcomes (Studnicka et al., 1997; Dockery et al., 1989; Braun-Fahrlander et al., 1997; McConnell et al., 1999, 2003; Shima and Adachi, 2000; Peters et al., 1999b). There does, however, appear to be an association between long-term exposure to NO2 and decreased lung function growth with age in children, based upon studies in southern California (Peters et al., 1999a, 1999b; Gauderman et al., 2000, 2002), but NO2 was correlated with other motor vehicle related pollutants, again implicating motor vehicle derived emissions in these effects. A study of adults in Europe (Schindler et al., 1998) indicated a relationship between NO2 exposure and changes in lung function, as noted by FVC. Some European studies have provided indication that chronic exposure to NO2 is associated with increased risk of all cause mortality (Hoek et al., 2002; Nafstad et al., 2004; Filleul et al., 2005). These studies have suggested a specific association of NO2 with cardiopulmonary mortality as well. However, studies in the United States (Dockery et al., 1993; Pope et al., 2002) do not seem to provide similar evidence for such an effect of chronic NO2 exposure and all-cause mortality. In summary, given the available epidemiological evidence, it is not possible to provide an unequivocal conclusion regarding adverse health effects of NO2. There have been both positive and negative findings at various levels of NO2 exposure, and with various degrees of precision in measuring actual exposure levels. There does appear to be a relationship between exposure and increased mortality due to all cause, or due to cardiovascular and respiratory effects, although the effect estimate is generally reduced when adjustments are made for other pollutants, specifically PM and ozone. Thus, while short-term variations in NO2 appear to correlate with increased daily mortality, a definitive causal relationship cannot be concluded. Some results are also suggestive that an increase in acute respiratory illness, especially in younger children, may be associated with chronic exposure, although the extent of any such effect or excess risk is small. While there also appears to be an effect from NO2, which is independent of that from other related air pollutants, the exact extent of the contribution of NO2 is not always clear. However, the strongest association does appear to be in asthmatics, and this is either not modulated, or modulated to a lesser degree, by other pollutants. A number of studies indicate there to be a fairly strong association between NO2 and hospital admissions or emergency room visits for asthma in children where NO2 was the only pollutant associated, or where adjusting for other pollutants did not affect the

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association with NO2. Finally, there is fairly strong evidence that chronic exposure to NO2 adversely affects lung function growth in children; this could be reflected in reduced function as adults. Thus, while both acute and chronic exposure to NO2 has been associated with adverse health outcomes, it is often unclear as to whether there is an independent effect; but there is clear evidence for traffic-related air pollution, which contains NO2, having adverse health effects. 22.3.4

Health Effects––Toxicology

Toxicological studies can be helpful in providing biological plausibility for health outcomes noted in epidemiological studies. However, a significant fraction of the NO2 toxicological database involves experimental exposures to concentrations >5 ppm. While such studies may help elucidate mechanisms of toxicity that influence responses to high concentrations, they are often of limited use in attempts to determine the public health significance from actual, much lower concentration, ambient air exposures. Thus, in this chapter, generally only studies using 5 ppm will be discussed. However, when necessary to help elucidate certain mechanisms, effects at higher levels will be presented as well. 22.3.4.1 Studies in Animals The largest database concerning the biological effects of NO2 is that derived from controlled exposures of laboratory animals. Since the mechanisms underlying many responses are similar across species, effects in these animals may have implications for humans. It should be borne in mind, however, that the exposure concentrations needed for comparable response likely differ between species. In any case, nitrogen oxides have been shown to exert a wide range of biological effects, mostly within the respiratory tract. Respiratory Tract Defenses Mucociliary transport provides a first line of defense against prolonged retention of deposited particles in the tracheobronchial tree. Acute (1–2 h) exposures to NO2 at levels 10 ppm did not alter mucociliary transport rate from the tracheobronchial tree of laboratory animals (Schlesinger, 1989). Rats exposed for 6 weeks to 6 ppm NO2 showed a transient depression in mucociliary activity (Giordano and Morrow, 1972), while rabbits exposed for 2 h/day for 14 days to 0.3 or 1 ppm did not show altered tracheobronchial mucociliary transport (Schlesinger et al., 1987). Thus, the available data suggests that with either single or short-term repeated exposures, higher than ambient levels are needed to alter tracheobronchial mucociliary transport. While the mechanisms underlying any such changes are not certain, they may involve NO2-induced effects upon ion transport across the airway epithelium (Robison and Kim, 1995) or upon ciliary beat activity (Ohashi et al., 1993). Particle clearance from the respiratory region of the lungs has also been assessed following exposures to NO2. Rats exposed to 1, 15, and 24 ppm showed a decrease in clearance after 22 daily exposures to 15 and 24 ppm, but accelerated clearance after exposures to 1 ppm (Ferin and Leach, 1977). Rabbits exposed for 2 h to 0.3, 1, 3, or 10 ppm showed accelerated clearance at all concentrations, while repeated 2 h/day exposures for 14 days to 1 or 10 ppm NO2 resulted in clearance patterns similar to those with single exposures at the same concentration (Vollmuth et al., 1986). Ferrets exposed to either 0.5 or 10 ppm NO2 for 4 h/day, 5 days/week for 8 or 15 weeks showed a reduction in clearance measured 12 weeks after the start of either exposure regime (Rasmussen et al., 1994).

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Alveolar macrophages play a central role in the defense of the lungs, and alterations in numbers and functional properties of these cells may affect susceptibility to disease or injury. Macrophage numbers increased with continuous exposure of rats to 17 ppm, but not with 2 ppm (Stephens et al., 1972), and after 7 days of continuous exposure of rats to 4 ppm (Mochitate et al., 1986). However, no change in cell number was found following exposure of rabbits to 0.3 or 1 ppm NO2 for 2 h/day for 13 days (Schlesinger, 1987). A subpleural accumulation of alveolar macrophages was found in rats exposed for 7 h/day, 5 days/week for 15 weeks to 5 ppm NO2, but not to 1 ppm (Gregory et al., 1983). Rombout et al. (1986) noted some increase, by 2 days, in the number of macrophages in terminal bronchioles and adjacent alveoli in the lungs of rats exposed continuously to 5 ppm NO2; this was not seen with 1 or 2.5 ppm. Others have noted concentration-related increases in macrophage numbers with exposure to 5–40 ppm for 2 days to 15 week (Kleinerman et al., 1982; DeNicola et al., 1981; Busey et al., 1974; Wright et al., 1982; Foster et al., 1985). Various functional properties of macrophages essential to adequate defense, for example, surface attachment, mobility, and phagocytosis, have been assessed following exposure to NO2. Schlesinger (1987) exposed rabbits to 0.3 or 1 ppm for 2 h/day and found no effect on attachment, but a depression of mobility at day 3 in the 0.3 ppm group. Macrophages obtained from baboons exposed to 2 ppm for 8 h/day, 5 days/week for 6 months showed reduced responsiveness to migration inhibitory factor, a lymphokine that mediates cell movement (Greene and Schneider, 1978). Suzuki et al. (1986) found depressed phagocytic activity in macrophages obtained from rats exposed for 10 days to 4 or 8 ppm NO2, while Lefkowitz et al. (1986) noted no change in such activity in macrophages from mice exposed for 7 days to 5 ppm. The phagocytic activity of rabbit macrophages was reduced by in vivo exposure to 0.3 ppm, but was enhanced with exposure to 1 ppm by 3 days, and returned to control values by 7 days and remained there through 13 days of exposure (Schlesinger, 1987). An exposure–concentration dependent difference in the direction of phagocytic response seems to be a characteristic of NO2. Thus, Schlesinger (1989) found a reduction in phagocytic activity of macrophages recovered immediately after a 2-h exposure of rabbits to 1 ppm NO2; with 10 ppm, no change was seen immediately after exposure, but activity was increased 24 h postexposure. Ehrlich et al. (1979) found exposure of mice to 0.5 ppm NO2 for 3 h/day, 5 days/week for 2 months to depress phagocytosis, while Sone et al. (1983) showed enhanced phagocytosis in macrophages obtained from rats exposed to 40 ppm for 4 h/day for 7 days. The reasons for such differences in direction of response are unknown. Macrophages are a source of various biological mediators, and their ability to produce these may be compromised by pollutant exposure. The eicosanoids are a class of mediators produced in response to a wide variety of cellular perturbation and have various effects on airway physiology and the immune system. Alveolar macrophages obtained from rats exposed to 0.5 ppm NO2 for 0.5–10 days exhibited complex responses related to the production of eicosanoids (Robison et al., 1993). An initial depression of production was followed by recovery for some of these mediators, but not others, with increasing exposure duration. The complexity of response was also noted in a study of rat alveolar macrophages acutely exposed in vitro to 0.1–20 ppm NO2 (Robison and Forman, 1993). Low concentrations (up to 5 ppm) had small effects on basal synthesis of eicosanoids but amplified response to stimulated production of eicosanoids, while high concentrations (20 ppm) showed the reverse pattern of response. Finally, rats continually exposed to 10 pm for 1, 3, or 20 days showed reduced levels of tumor necrosis factor in lavage and suppression of cytokine

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signaling-3 mRNA, both of which were likely due to NO2-induced changes in activation state of the macrophages (Garn et al., 2003). NO2 may impair the ability to resist infectious agents; this is suggested by some epidemiological studies noted above. Mice exposed continuously to 0.5 ppm NO2 showed increased mortality to K. pneumonia after 3 months of exposure (Ehrlich and Henry, 1968), while 0.05 ppm for 24 h/day for 15 days did not change bacterial resistance (Gardner et al., 1982). The finding of increased susceptibility does, however, depend upon the specific organism being used. Thus, while exposure of mice for 3 h/day for 3 months to 0.5 ppm increased mortality to Streptococcus sp. (Ehrlich et al., 1979), exposure to 0.5 to 1.5 ppm NO2 continuously for 3 months produced no effect on mortality due to K. pneumoniae (McGrath and Oyervides, 1985); on the contrary, exposure to 5 ppm for 3 days did result in enhanced mortality. It may be that peak exposure and exposure pattern are important modulators of response to NO2. A number of infectivity studies involved exposure to a baseline NO2 concentrations upon which spikes to a higher level were superimposed to mimic ambient exposures. The relative effect of such spikes is not always clear, but seems to depend upon both spike duration and time between spikes. Miller et al. (1987) noted that mortality due to infection was greater in a spike regimen (to 0.8 ppm) than in the baseline-exposed group (0.2 ppm). Others have found that both the number and amplitude of spikes are of importance in increasing mortality (Gardner et al., 1979; Graham et al., 1987). In fact, effects from such exposure excursions may approach those due to more continuous exposure to a lower concentration. This is consistent with the notion that, in general, brief exposures to high NO2 levels are more hazardous than are longer duration exposures to lower concentrations (Lehnert et al., 1994). The effect of NO2 on mortality due to bacterial infection appears to increase with both exposure duration (T) and peak concentration (C), although the latter seems to have more influence than the former for fixed C T values (Gardner et al., 1979). Any differences between intermittent and continuous exposure also seem to disappear as the number of days of exposure increases (Gardner et al., 1979). Other studies suggest that, as concentration increases, a shorter exposure time is needed for intermittent and continuous exposure regimes to produce similar degrees of effect (Ehrlich and Henry, 1968; Ehrlich, 1979). Mortality is also proportional to exposure duration if the bacterial challenge is given immediately after exposure, but may not be when the challenge is given much later (Gardner et al., 1982). For example, no effect of 3.5 ppm NO2 for 2 h was seen in mice when bacteria were administered 27 h after exposure, while increased mortality was evident when administration was immediately after NO2 inhalation (Ehrlich, 1980). Effects of 25 ppm for 2 h on mice were seen only when the microbial challenge was given within 72 h after NO2 exposure (Purvis and Ehrlich, 1963). These results suggest that a critical time frame exists between exposure and bacterial challenge after which NO2 will not affect resistance. The mechanism(s) underlying any NO2-induced change in host resistance to bacteria are not known. However, since exposure levels that alter resistance do not affect physical clearance processes, the response to NO2 may be due to impaired intracellular killing of microbes, perhaps reflecting macrophage dysfunction. For example, macrophages are a source of numerous biochemical mediators that are directly involved in antibacterial action, for example, superoxide anion, and a depression in superoxide production has been noted following NO2 exposure in some studies (Amoruso et al., 1981; Suzuki et al., 1986; Robison et al., 1993), although at higher than ambient levels. However, human alveolar macrophages exposed in vitro to 0.1–0.5 ppm NO2 for 30–120 min showed increased reactive oxygen

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intermediate production in a dose-dependent fashion (Kienast et al., 1994). The effects of NO2 on viral infectivity have also been examined. Exposures to 0.5 ppm or greater on a continuous basis likely increase susceptibility, while at higher concentrations the exposure duration needed for any effect is lowered (Ito, 1971; Rose et al., 1988). Furthermore, environmental stresses may enhance the lethality of infectious agents over and above that due solely to NO2 exposure. These may include exercise (Illing et al., 1980), elevated temperature (Gardner et al., 1982), and the presence of other pollutants. Exposure to NO2 may affect allergic response. Rats exposed to 5 ppm for 3 h after sensitization with house dust mite antigen had higher levels of serum IgE and local respiratory tract IgA, IgG, and IgE antibodies than did controls (Gilmour, 1995). The exposed animals also had increased lymphocyte activity in the spleen and local lymph nodes and showed an increase in respiratory tract inflammatory cells. This suggests that NO2 may enhance immune responsiveness and increase the severity of pulmonary inflammation in sensitized lungs and may, thus, play some role in the exacerbation of immune-mediated respiratory disease. A number of studies have examined the effects of NO2 on specific parameters of respiratory and/or systemic humoral and cellular immunity. While immune suppression and/or enhancement of factors involved in airway hyperreactivity has clearly been shown to follow exposure to levels of NO2 above 5 ppm, as evidenced by various end points including the response of T-cells, antibodies, or production of interferon or other inflammatory mediators (e.g., Holt et al., 1979; Valand et al., 1970; Fujimaki and Shimizu, 1981; Campbell and Hilsenroth, 1976; Ayyagari et al., 2004), there are only a few reports of response to lower levels. These studies suggest that short-term repeated exposures may result in reductions in counts of lymphocytes in the lungs or spleen, or a depression in antibody responsivity to particular antigens. Mice exposed for 7 h/day, 5 days/week for 7 weeks to 0.25 ppm NO2 showed reduced total T-lymphocyte numbers in the spleen, with concomitant reductions in certain subpopulations of these cells, for example, helper cells (Richters and Damji, 1988). Exposure of mice for 3 months to 0.5 ppm NO2 resulted in a depressed responsiveness of both T- and B-lymphocytes in spleen (Maigetter et al., 1978). No effect on the cell-mediated immune system was found either in mice exposed for 24 h to 5 ppm NO2 (Lefkowitz et al., 1986), or in those exposed to 1.6 ppm NO2 for 4 weeks (Fujimaki et al., 1982). Mice exposed to 0.4 or 1.6 ppm NO2 for 4 weeks showed depressed primary antibody responsivity to sheep red blood cells in vitro (Fujimaki et al., 1982) while mice exposed to 4 ppm NO2 continuously for up to 56 days showed no change in the antibody response to Tcell-dependent and independent antigens in spleen (Fujimaki, 1989). In another study, mice were vaccinated with influenza virus after they had undergone 3 months of continuous exposure to 0.5 or 2 ppm NO2 with daily spikes (1 h) of 2 ppm for 5 days/week. Both concentrations resulted in a reduction in mean serum neutralizing antibody titers (Ehrlich et al., 1975). Guinea pigs exposed to 1 ppm for 6 months showed a reduction in all immunoglobulin fractions (Kosmider et al., 1973). On the contrary, Balchum et al. (1965) noted an increase in serum antibody titers against lung tissue in guinea pigs exposed to 5 ppm NO2 for 4 h/day after 160 h of exposure, and further increases as exposure duration increased. Enhanced immune function may be just as detrimental as suppressed function, through overstimulation of response and hypersensitivity. As with other end points, the effects of NO2 upon the immune system appear to be related to various exposure parameters. While some studies show no effects, others show enhancement or depression of immune parameters, depending upon the exposure concentration, the

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length of exposure, and the animal species used. In addition, the direction of change appears to depend upon exposure concentration. For example, humoral response in monkeys chronically exposed to NO2 was enhanced at a low concentration (1 ppm), but suppressed at a higher level (5 ppm) (Fenters et al., 1971, 1973). Respiratory Tract Structure Exposure to NO2 may produce structural alterations in the respiratory tract. As noted, the anatomic region most sensitive to NO2 is the area encompassing the terminal and respiratory bronchioles and adjacent alveolar ducts and alveoli. The primary cellular targets within this region are ciliated cells of the bronchiolar epithelium and Type 1 cells of the alveolar epithelium. Acute exposure to NO2 can result in hypertrophy and hyperplasia of alveolar Type 1 cells, followed by cell death and desquamation and proliferation of and replacement by Type 2 cells. The end result can be a thickened air–blood barrier. The bronchiolar response is characterized by hypertrophy and hyperplasia of epithelial cells, loss of secretory granules and surface protrusions of Clara cells, and loss of ciliated cells, or of cilia. With chronic exposure, many of these same changes are seen, but there is increased cilia loss over larger areas of epithelium and in more proximal airways, and the structure of the remaining cilia may be altered. The temporal progression of NO2-induced lesions has best been described for the rat. The earliest alterations resulting from concentrations 2 ppm occur within 24–72 h of continuous exposure, with repair of injured tissue and replacement of destroyed cells beginning within 24–48 h of continuous exposure. Division of Type 2 cells is observed within 12 h after initial NO2 exposure, the rate becoming maximal by about 48 h, and then decreasing to preexposure levels by about 6 days, even with continued exposure. In some cases, the resolution of NO2-induced morphologic changes may be complete after exposures end; on the contrary, some lesions may resolve while others remain, even when exposure continues (Rombout et al., 1986; DeNicola et al., 1981; Kubota et al., 1987). Chronic exposure to NO2 may result in alterations in lung architecture resembling emphysema-like disease, for example, enlargement of airspaces, increase in mean linear intercept (a measure of the distance between alveolar walls), and reduction in the internal surface area of the alveolar region. However, the relationship between exposure and the development of emphysema remains unclear. A problem in evaluating reported emphysematic changes in animal models is the definition of the disease, which has changed over the years and which has been defined differently by various professional groups (NIH, 1985). While long-term exposure to high NO2 concentrations (>10 ppm) are required to produce clearly definable emphysema-like changes (e.g., Barth et al., 1995), there is evidence that lower NO2 levels may result in emphysema, emphysema-like changes, or altered alveolar dimensions if present in complex mixtures of NOx (Hyde et al., 1978) or when administered during lung development (Rasmussen and McClure, 1992). However, clear evidence of changes characteristic of human emphysema, that is, alveolar septal degeneration, enlarged airspaces, and associated functional changes, is absent with exposure at low levels. There is, however, some evidence for changes similar to those seen in human emphysema with exposure to high concentrations. These involved exposures to levels ranging from 8 to 20 ppm for up to 33 months (Haydon et al., 1967; Freeman et al., 1972). Another study that involved exposure of dogs for 5.5 years to a mixture of NO2 at 0.64 ppm and NO at 0.25 ppm followed by a postexposure period of 2.5 years noted structural changes similar to human centrilobular emphysema that were noted after the postexposure period had ended (Hyde et al., 1978).

836

NITROGEN OXIDES

While the extent and degree of structural alterations induced by NO2 appear to be related to exposure concentration, little is known about effects of other modifying factors, for example, exposure duration or the temporal pattern of exposure. The contribution of exposure time in the histopathologic response to acute inhalation was examined in rats (Stavert and Lehnert, 1988). The most pronounced effects were found with the highest NO2 concentration in any particular set of exposures where the product of concentration and time was equivalent, indicating that concentration played a more important role than did exposure time in tissue injury. This is consistent with the relative roles of C and T in infectivity, discussed previously. Another study (Rombout et al., 1986) assessed the concentration–time response relation for intermittent and continuous exposures, and likewise concluded that concentration played a more important role in inducing morphologic lesions than did exposure duration, as long as the product of C T was constant. The effect of concentration was found to be greater with intermittent than with continuous exposure, and the onset of response was also delayed with intermittent compared to continuous exposure. The morphological effects of exposure patterns involving transient spikes were examined in a number of studies (Gregory et al., 1983; Miller et al., 1987; Crapo et al., 1984; Chang et al., 1986). Results are equivocal, and it is not clear whether these peaks significantly contributed to morphological damage in excess of that due to integrated exposure. In spite of the fact that there is a fairly extensive database concerning morphologic effects of NO2 in animal models, it is still quite difficult to establish a threshold exposure condition for these end points. This is due to the great complexity of changes occurring with exposure, as well as to large interspecies differences in response. For example, the rat appears to be less sensitive to NO2 compared to other species, such as the guinea pig or monkey. Furthermore, different cell types show differential sensitivity to NO2. In general, morphological alterations, some of which may be persistent, are found with chronic exposure to concentrations 25 ppm. A study involving long-term exposures (Kobayashi and Miura, 1995) involved exposure of guinea pigs to 0.06, 0.5, 1, 2, or 4 ppm NO2 for 24 h/day for 6 or 12 weeks. Airway responsiveness to histamine and specific airway resistance were assessed on the last day of each exposure. Exposure to 2 and 4 ppm by 6 weeks of exposure resulted in increased airway responsiveness, while exposure to the same concentration resulted in increased resistance by 12 weeks of exposure. Tepper et al. (1993) performed a long-term exposure of rats to NO2. Animals were exposed to NO2 having a 0.5 ppm background with 1.5 ppm peaks (2 h) for up to 78 weeks. No exposure-related changes in nitrogen washout, compliance, lung volume, or CO diffusion capacity were noted, but at 78 weeks there was some reduction in a measure of forced expiratory flowrate. However, the authors indicated that the change was borderline, and suggested that long-term exposure to high ambient urban levels did not lead to any dysfunction suggestive of degenerative lung disease. The overall database suggests that NO2 at realistic levels in terms of ambient exposure has not been shown to significantly alter pulmonary mechanics or bronchial responsivity in animal models, consistent with results of controlled clinical studies in humans. Extrapulmonary Effects Exposure to NO2 may affect target sites beyond the respiratory tract. End points that have been shown to be altered include body weight, blood cell counts, blood cell membrane and serum chemistry, liver and kidney function, brain protein enzymes, and neuromotor function (e.g., Graham et al., 1982; Tabacova et al., 1985; Freeman et al., 1966; Wagner et al., 1965; Case et al., 1979; Kaya et al., 1980; Kaya and Miura, 1982; Mochitate et al., 1984; Kosmider et al., 1973; Miller et al., 1980; Takahashi et al., 1986; Sherwin and Layfield, 1974). However, the data are conflicting and, because of this, the ability to relate reported changes to human health effects is severely limited.

NITROGEN DIOXIDE

839

Carcinogenicity/Reproductive Toxicity Exposure to NO2 even at high levels does not seem to be genotoxic or teratogenic in appropriate assay systems (Kripke and Sherwin, 1984; Gooch et al., 1977). While one study did note an increase in the rate of DNA strand breaks in hamster cells exposed in vitro to 10 ppm for 20 min, exposure to 5 ppm for up to 30 min had no such effect (G€ orsdorf et al., 1990), and in vivo exposure of mice to 20 ppm for up to 23 h did not result in any genotoxicity (Victorin et al., 1990). This apparent conflict in response may be due to repair mechanisms operating in vivo that are not operative in in vitro assays. The ability of NO2 to act as a carcinogen, or cocarcinogen, is unclear, but there is no direct evidence that NO2 exposure results in the development of tumors. Some concern is, however, based upon the fact that exposure can result in nitrite in blood and this, in turn, may produce carcinogens, such as nitrosamines, after further reaction in the body. Although there have been no long-term carcinogenesis bioassays performed with NO2, one chronic inhalation study, in which mice were exposed to 1, 5, and 10 ppm NO2 for 6 h/day, 5 days/week for 6 months, suggested a small increase in tumor (pulmonary adenoma) frequency and incidence in the highest dose group (Adkins et al., 1986). However, such data must be interpreted with caution, and the relationship between cancer development in mice and that in humans is not clear. Although not likely a carcinogen itself, NO2 may modulate tumorogenic processes in the lungs (Witschi, 1988). For example, in conjunction with a specific carcinogen, NO2 exposure may be involved in the pathogenesis of small cell carcinoma (Witschi, 1988), especially since it has been shown to modulate the number of neuroendocrine cells, the precursor cells for this disease (Kleinerman et al., 1981; Palisano and Kleinerman, 1980). As another example, an enhancement of tumor colonization in the lungs of mice injected (IV) with melanoma cells was noted after exposure to NO2 at 0.4 or 0.8 ppm for 8 h/day, 5 days/week for 10–12 weeks (Richters and Kuraitis, 1981). This could be due to injury of lung capillary endothelium by NO2, facilitating metastases of blood-borne cancer cells to the lungs (Richters and Richters, 1989), or to the suppression of immune system components. However, as with other end points, the database regarding the role of NO2 in carcinogenic processes is conflicting. For example, NO2 has been shown to actually enhance the cytotoxic response of macrophages (Sone et al., 1983), which implies greater antitumor defense capabilities. Thus, any role for NO2 in cancer etiology requires further evaluation. Is there any epidemiological evidence that NO2 is involved in the etiology of cancer? Some studies do suggest a relationship between NO2 and lung cancer (e.g., Hoek et al., 2002; Nafstad et al., 2004). However, NO2 is generally associated with other pollutants from the same source, many of which are known carcinogens, so all results related directly to NO2 must be interpreted with great caution. There have been some epidemiological studies suggesting that exposures to mixtures containing NO2 during pregnancy may be associated with fetal/reproductive effects, such as low birth weight and perinatal mortality but, again, any independent effect from NO2 is unclear (Liu et al., 2003; Wilhelm and Ritz, 2003). 22.3.4.2 Controlled Studies with Humans By their nature, studies with human volunteers can only be used to evaluate transient effects of acute exposure. They have generally used changes in standard respiratory mechanical indices as markers of response; a few studies, however, have employed other end points, which include bronchoprovocation challenge testing, clearance of inhaled aerosols, and analysis of biochemical and cellular

840

NITROGEN OXIDES

components of bronchopulmonary lavage. Various subject groups have been examined. These include healthy individuals with no history of respiratory disease, allergy, etc., as well as people with allergies or a history of asthma or chronic obstructive pulmonary disease (COPD). Acute exposures (up to about 2 h) to NO2 at levels 1 ppm have not resulted in any consistent, significant changes in respiratory mechanics in normal, healthy adult subjects at rest (Beil and Ulmer, 1976; Bylin et al., 1985; Hazuch et al., 1982; Koenig et al., 1985; Bascom et al., 1996). Regarding higher levels, the study of Beil and Ulmer (1976) involving 2-h exposures to 1, 2.5, 5, and 7.5 ppm, indicated a change in total respiratory resistance occurring at 2.5 ppm, although the effects were quite small. von Nieding et al. (1973) noted a decrease in diffusing capacity (DLco) with a 15-min exposure to 5 ppm. Finally, no changes in lung mechanical function were found with exposure to 2 ppm for 2 h, or for 2 h/day for 3 days (Mohsenin, 1988; Goings et al., 1989). Exposure to 1 ppm NO2 in conjunction with various degrees of exercise have also resulted in inconsistent effects on respiratory mechanics in healthy people; most of the studies showed no effects which could be unequivocally attributed to NO2 (Folinsbee et al., 1978; Frampton et al., 1989a; Hackney et al., 1978; Kerr et al., 1979; Morrow and Utell, 1989). Reduced compliance was noted following exposure for 2 h at 0.5 ppm (Kulle, 1982), but the meaning of this with a lack of other lung mechanical changes was not clear. Linn et al. (1985a) found no change in resistance or spirometry with exposure at 4 ppm for 75 minutes. Increased airway responsivity in healthy subjects has been noted following a 2-h exposure to 7.5 ppm (Beil and Ulmer, 1976), with 1 h to 2 ppm (Mohsenin, 1988), and with 3 h (with intermittent exercise) to 1.5 ppm (Frampton et al., 1989a). Again, however, the results are not consistent, with other studies at similar concentrations and exposure durations finding no change (e.g., Kulle and Clements, 1987). Exposures at 0.6 ppm have not produced any change in responsivity at all (Morrow and Utell, 1989; Frampton et al., 1989a; Bylin et al., 1985; Hazuch et al., 1983). Particular subsegments of the population may be especially susceptible to the effects of NO2. As noted in epidemiological studies, one such group is asthmatics. A number of studies have been performed with exposure levels ranging from 0.1 to 4 ppm for durations ranging up to 4 h, usually with exercise; effects on various aspects of mechanical function, such as spirometry or airway resistance, have ranged from none to slight, and all with much inconsistency (e.g., Avol et al., 1988; Bauer et al., 1986; Ahmed et al., 1982; Hazuch et al., 1982, 1983; Bylin et al., 1985; Koenig et al., 1985, 1987; Kerr et al., 1979; Rubinstein et al., 1990; Morrow and Utell, 1989; Roger et al., 1990; Kleinman et al., 1983; Linn et al., 1985a; Mohsenin, 1987; Salome et al., 1996; Morrow et al., 1992). A study that measured pulmonary function in adult asthmatics in their home and also monitored indoor NO2 levels noted that average exposures to >0.3 ppm produced a decline in certain pulmonary function measures, but inconsistent effects were seen at lower exposure levels (Goldstein et al., 1988). Finally, when there is any response, it may only occur with exercise. Exposure to 0.3 ppm for 30 min produced no change in pulmonary mechanics indices in resting asthmatics, but effects were noted when exercise was incorporated into the exposure protocol (Bauer et al., 1986). The most sensitive pulmonary mechanical response to NO2 in people with airway disease appears to involve changes in airway responsiveness. However, there is much variabilty in results from different studies, and also an apparent lack of a dose–response relationship. While some studies have indicated increased responsiveness due to NO2 exposures at 0.14– 0.5 ppm (Bauer et al., 1986; Bylin et al., 1988; Kleinman et al., 1983; Mohsenin, 1987;

NITROGEN DIOXIDE

841

Salome et al., 1996; Strand et al., 1996), others have indicated no such effects at similar levels (Hazuch et al., 1983; Roger et al., 1990; Linn et al., 1986; Orehek et al., 1981; Avol et al., 1988; Bylin et al., 1988), and exposure to a much higher level (3 ppm) has also produced no effect (Linn et al., 1986). Any NO2-induced increased responsiveness may occur with a several-hour delay following exposure in asthmatics (Strand et al., 1996). Potentiation of cold-induced airway constriction and airway responsiveness to histamine in asthmatics was enhanced by exposure to 0.3 or 0.26 ppm, respectively (Bauer et al., 1986; Strand et al., 1996). A meta-analysis of a number of studies involving both asthmatics and normals indicated that acute exposure to NO2 would enhance responsiveness to various stimuli with exposure to at least 0.11 ppm in asthmatics, but at least 1 ppm in normals (Folinsbee, 1992). While the mechanism of any NO2-induced hyperresponsiveness is not known, it may involve alterations in the metabolism of endogenous bronchoconstrictors (Hoshi et al., 1996) or activation of specific cells within the airways (Ohashi et al., 1993). Mild asthmatics and normals were exposed to 1 ppm NO2 (with intermittent exercise) for 3 h, followed by bronchopulmonary lavage 1 h postexposure. While no change in differential cell counts was noted in either group, the asthmatics showed changes in lung eicosanoids not seen in normals, suggesting that NO2 could activate cells compatible with airway inflammation (Jorres et al., 1995). Even if asthmatics or allergic individuals may not show any enhanced response directly to NO2, exposure may alter their response to antigens. A number of studies have examined response to NO2 in terms of enhancement of the response to inhaled allergens in sensitized individuals. Humans having a history of allergic rhinitis were exposed to 0.4 ppm NO2 for 6 h, followed by challenge with an allergen (Wang et al., 1995). There was some evidence that NO2 primed eosinophils for subsequent activation by the allergen. Acute exposure to 0.43 ppm NO2 enhanced airway constriction in mild asthmatics in response to inhaled house dust mite antigen (Tunnicliffe et al., 1994). Similarly, airway constriction to pollen was enhanced in allergic asthmatics acutely exposed to 0.27 ppm NO2 (Strand et al., 1997) and following repeated exposure at 0.27 ppm (Strand et al., 1998). Finally, allergic asthmatics were exposed in a roadway tunnel to NO2 at a median level of 0.17 ppm, but ranging from 0.11 to 0.25 ppm, for 30 min; subsequent inhalation of an allergen resulted in greater early asthmatic reaction and more symptoms during the later phase asthmatic response when compared to air control exposed individuals (Svartengren et al., 2000). However, one must realize that roadway exposure involves more than just NO2, so effects may not have been due to NO2 alone, or at all. Nitrogen dioxide induced modulation of response to antigens may be due to recruitment of eosinophils (Barck et al., 2002). Thus, for example, subjects with allergic asthma were exposed to 0.27 ppm for 15 min on one day and for two 15-min intervals the next day; they were noted to have increased levels of eosinophil cationic protein, a component found in eosinophil granules, in both systemic blood and sputum when subsequently exposed to allergen. Another possibly sensitive subsegment of the population is people with COPD, that is, chronic bronchitis and emphysema. Increased airway resistance has been found in individuals with COPD after exposure to 1.6 ppm in conjunction with exercise (von Nieding and Wagner, 1979), while a decrease in FVC was noted following exposure to 0.3 ppm for 4 h with intermittent exercise (Morrow and Utell, 1989), and a decrease in FEV1 was noted following exposure for 1 h to 0.3 ppm (Vagaggini et al., 1996). On the contrary, no changes in airway resistance in chronic bronchitics exposed to 0.5 ppm for 2 h with exercise, or in

842

NITROGEN OXIDES

spirometry of COPD patients exposed to 0.5–2 ppm for 1 h also with exercise, have been noted (Kerr et al., 1979; Linn et al., 1985b). Thus, the database is currently not sufficiently robust to allow determination of the specific exposure conditions, that is, concentration, duration, and ventilation, for threshold effects on lung function in healthy humans with acute exposure. Lung mechanics may, in fact, not provide very sensitive indices of response in such people. On the contrary, functional changes may occur in individuals with asthma and/or COPD following exposure to lower levels of NO2 than those affecting normals. Again, however, the results are inconsistent. Results of one study examining pulmonary functional indices with asthmatics have not been confirmed by a subsequent one, or responses of a particular subject group are not always reproducible (Orehek et al., 1976; Hazuch et al., 1983; Bauer et al., 1985; Bromberg, 1988). There is, however, some evidence that especially sensitive subgroup(s) may exist within the asthmatic population (Bauer et al., 1986; Morrow and Utell, 1989). That is, the variability in responses noted above may be the result of differences in the severity or type of asthma in the subjects examined within one study, or between different studies. Asthmatics also exhibit a wide range of response to external stimuli, so some variability may merely be due to an interindividual variation in response to NO2. The lowest concentration that does result in observed effects on airway responsivity in exercising asthmatics is in the 0.2–0.5 ppm range; in normals, levels of 5 ppm may cause bronchoconstriction, but minimum levels of at least 1– 2 ppm are generally needed for changes in pulmonary functional parameters. Most mild asthmatics are not sensitive to NO2 at less than or equal to 0.6 ppm, at least in terms of changes in respiratory mechanics, while nonspecific airway responsiveness in mild asthmatics may be increased at levels >0.1 ppm. Controlled clinical studies have examined other aspects of pulmonary biology after exposure to NO2. Humans exposed for 20 min to 1.5–3.5 ppm NO2 did show a reduction of mucociliary activity measured 45 min following exposure (Helleday et al., 1995). The effects upon infectivity of an attenuated influenza virus in healthy humans was assessed by Kulle and Clements (1987); NO2 exposure levels were 1 to 3 ppm. There were no overall statistically significant changes in infectivity rates, although they were elevated in some of the NO2 exposed groups. In another study (Goings et al., 1989), there was suggestive evidence that exposure for 2 h/day for 3 days to 1 or 2 ppm NO2 increased susceptibility to respiratory viruses in healthy adults. Frampton et al. (1989b) examined the effect of NO2 exposure in vivo on the ability of alveolar macrophages to inactivate influenza virus in vitro. Healthy humans were exposed either to 0.6 ppm for 3 h or to 0.05 ppm for 3 h with three 15-min spikes to 2 ppm. There appeared to be less effective inactivation of the virus by macrophages harvested from the humans exposed to 0.6 ppm, but the results just missed statistical significance. No effects were noted in the individuals exposed to the lower concentration with the 2 ppm spikes. There also seemed to be a trend of increased production of interleukin-1 (IL-1) by macrophages from some individuals, namely those whose cells tended to have reduced viral inactivation activity. Effects on IL-1 were also examined by Pinkston et al. (1988), with exposure of macrophages harvested by lavage to 5–15 ppm NO2 for 3 h. No change in cell viability nor in release of IL-1 was noted. In any case, increased infectivity in NO2-exposed laboratory animals together with the above suggestive findings in humans indicates that NO2 may indeed alter host defense in humans. Healthy subjects exposed to 1–3 ppm NO2 for 2hr/day for 3 days and then exposed to attenuated influenza virus showed a slight trend toward increased infectivity (Goings et al., 1989). However, exposure for 3.5 h to 0.6 ppm NO2 resulted in a decreased inactivation of the influenza virus by alveolar macrophages (Frampton et al., 1989b).

NITROGEN DIOXIDE

843

Healthy subjects exposed to 0.6 ppm for 2 h on 4 days showed a small increase in the percentage of NK-lymphocytes (Rubinstein et al., 1991), but repeated exposures to 1.5 or 4 ppm for 20 min every other day for a total of 6 days reduced numbers of both B- and NKlymphocytes and altered the ratio of CD4 þ /CD8 þ cells (Sandstr€ om et al., 1992a, 1992b). Healthy subjects exposed to 1.92 ppm NO2 for 4 h on 4 days showed upregulation of the expression of IL-5, IL-10, IL-13, and ICAM-1 (Pathmanathan et al., 2003); the effects on the interleukins suggest that repeated exposure may exert a proallergic effect on the airway epithelium, while the effect on ICAM suggests a mechanism for neutrophil influx into the epithelium during an inflammatory response. Some other biochemical effects of inhaled NO2 have been examined in controlled clinical studies. In vitro exposure of human blood to high levels (>6 ppm) of NO2 has been shown to result in production of methemoglobin (metHb) (Chiodi et al., 1983), but Chaney et al. (1981) found no such change in normal humans exposed for 2 h to 0.2 ppm. A reported elevation of glutathione in these exposed subjects was not supported by the results of Posin et al. (1978), who found no effect following exposure to 1 ppm. The results of some of these studies are clouded by the lack of any consistent dose–response relationship. Exposure to 4 ppm NO2 for 20 min resulted in an inflammatory response in healthy individuals, as evidenced by changes in lymphocyte counts in lavage fluid obtained 4–24 h after exposure (Sandstr€ om et al., 1990). This is, however, not a consistent finding in humans (e.g., Mohsenin and Gee, 1987), possibly due to differences in experimental protocols, such as the times at which lavage was performed after exposure. Thus, exposure to 0.3 ppm for 1 h with exercise produced no acute inflammation in the proximal airways of normals, asthmatics or people with COPD (Vagaggini et al., 1996). It should be noted that exposures of laboratory animals to NO2 at levels up to 8 ppm for up to 10 days did not produce evidence of acute inflammation (Schlesinger et al., 1987; Gregory et al., 1983; Mochitate et al., 1986; Suzuki et al., 1986). Perhaps NO2 is not very effective in eliciting an inflammatory response at ambient levels with short-term exposure. Nitrogen dioxide exposure has been associated with development of emphysema in animal models. A component of the lungs’ defense against proteolysis is a-1-protease inhibitor. Mohsenin and Gee (1987) noted a decrease in levels of this enzyme in the lavage fluid of subjects exposed to 3–4 ppm for 3 h. However, the investigators noted that the extent of the decrease was not associated with any increased risk of emphysema. On the contrary, exposure of normal humans for 3 h (with intermittent exercise) to 1.5 ppm, or for 3 h to 0.05 ppm with three 2 ppm peaks, did not result in any change in activity of a-1-protease inhibitor in lavage fluid (Johnson et al., 1990). A 3-h exposure to 0.6 ppm resulted in an increase in levels of another antiprotease (a-2-macroglobulin) in lung lavage (Frampton et al., 1989c). In another study of potential lung damage, normal humans exposed to 0.6 ppm NO2 for 4 h/day for 3 days showed no effect on the excretion of hydroxyproline, a marker for connective tissue injury (Muelenaer et al., 1987). Effects of repeated exposures, which would more likely be involved in disease development, on these end points are unknown. 22.3.5

Health Effects––Summary and Conclusions

A large database exists concerning biological responses resulting from the inhalation of NO2. While there have been a significant number of epidemiology studies conducted over the past 10 years, there are very few new toxicology studies aimed at assessing mechanisms of response to NO2. In any case, comparisons between animal studies, controlled human exposures or epidemiologic studies is difficult, since the assays used in these different types

844

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of evaluations are not always directly comparable. One type of response index that has been examined in all of these studies is respiratory mechanics. However, changes in pulmonary function may not be very sensitive to NO2 due to the tendency of such tests to reflect changes in the large airways while the major targets for NO2 are the smaller conducting airways and respiratory region. In any case, there is little evidence that exposure of normal humans or laboratory animals to 1 ppm NO2 affects standard pulmonary mechanics responses. Even exposure to higher levels has resulted in inconsistent results. Airway responsiveness may be increased in normal human subjects, but generally only with exposures at >1 ppm NO2. Epidemiological data suggest that there may be long-term effects of NO2 on pulmonary function in children. Asthmatics may represent a population subgroup showing susceptibility to NO2. However, even among asthmatics, responses were not always consistent or reproducible, and those that have occurred involved increased airway responsiveness rather than changes in standard respiratory function indices. Surprisingly, effects noted in some studies at 1 ppm NO2 have not always been found with higher levels (up to 3 ppm), and this apparent lack of dose–response complicates any evaluation of the health significance of NO2 exposure. While it is possible that differences in the degree of asthma severity in the subjects used in the various studies may have accounted for some of this discrepancy, it does seem that mild asthmatics are not generally sensitive to NO2 concentrations 24 h

Chest pain

None

Duration of response

None

(c) Impact of various functional and/or symptomatic responses Interference with normal activity

Small

Small functional Normal and/or mild functionaland/or symptomatic symptomatic responses responses

Large functional Moderate and/or severe functional and/or symptomatic symptomatic responses responses

None

Many sensitive A few sensitive individuals individuals likely to limit activity likely to limit activity

Source: U.S. EPA (1996b).

None

872

OZONE

TABLE 23.2 (a–c) Gradation of Individual Responses to Short-Term Ozone Exposure in Persons with Impaired Respiratory Systems Functional Response

None

(a) FEV1 change

Decrements of 10% but 24 h

Normal

Mild

Moderate

Severe

None

With otherwise With shortness of normal breathing breath

Cough

Infrequent cough

Cough with deep breath

Chest pain

None

Duration of response

None

Discomfort just noticeable on exercise or deep breath 4 but 24 h

>24 h

(c) Large functional Moderate Small functional Impact of various Normal and/or severe functional and/or and/or mild functional functional and/ symptomatic symptomatic symptomatic or symptomatic and/or responses responses responses symptomatic responses sresponses Interference with None Few individuals Many individuals Most individuals normal activity likely to limit likely to limit likely to limit activity activity activity Medical No change Normal medication Increased frequency Increased treatment/self as needed or additional likelihood of medication medication use physician or ER visit Source: U.S. EPA (1996b).

BACKGROUND ON EXPOSURES AND HEALTH-RELATED EFFECTS

873

With respect to adversity, the 1996 staff paper concluded that responses listed as large or severe were clearly adverse. For responses listed as moderate, it was concluded that they could be considered adverse if there were repetitive exposures. Although we know a great deal about the transient effects following single exposure to O3 resulting from controlled laboratory exposures, and short-term responses in populations associated with peak ambient air concentrations, our current knowledge about the chronic health effects of O3 is much less complete. As discussed in the latter part of this chapter, the known chronic effects include alterations in lung function or structure. Such effects may result from cumulative damage and/or from the side effects of adaptive responses to repetitive daily or intermittent exposures. This review does not discuss the effects of O3 or its metabolites on nonpulmonary tissues or organs. It also does not discuss the health effects of increased ultraviolet radiation resulting from the depletion of stratospheric O3, which are viewed as minor in the 2006 O3 criteria document (U.S. EPA, 2006). This chapter provides a critical review of the health effects data and their significance to public health in relation to the populations exposed. The judgments made have been influenced by my participation in public CASAC reviews of U.S. Environmental Protection Agency (EPA) documents, but they differ, in some cases, from those of the EPA and of others on the CASAC panels.

23.2 BACKGROUND ON EXPOSURES AND HEALTH-RELATED EFFECTS 23.2.1

Sources and Distribution of O3 in Ambient Air

O3 in ambient air is attributable to several different sources. One is the intrusion of stratospheric O3, especially in the spring when the stratospheric–tropospheric air exchange is greatest. The other sources are driven by complex photochemical reaction sequences requiring input of HCs, NOx, and actinic radiation. Reactive organic vapors such as olefinic hydrocarbons, formaldehyde, and m-xylene, which are largely products of anthropogenic activities, are highly efficient contributors to O3 formation. On the other hand, methane (CH4), a major product of natural biogenic decay, and a relatively nonreactive hydrocarbon, can also contribute to O3 formation. Actually, the background concentration of CH4 has been rising over the last 100 years as a result of increasingly intensive agriculture and animal husbandry. The coincident increase in continental background O3 over the past century, from 10 to 20 ppb (Altshuller, 1987) to the current level near 40 ppb, may be due to the rising background of both CH4 and NOx. The NOx concentrations have grown continuously as fossil fuel usage has increased. The increase in NOx may also account for a greater rate of O3 formation by photochemical reactions with isoprene and terpenes emitted by trees. As also noted by Altshuller (1987), the role of NOx in tropospheric O3 formation is especially critical. Unless NOx concentrations exceed about 0.02 to 0.03 ppb, photochemical O3 loss exceeds photochemical O3 production. There are remote regions of the troposphere where the NOx concentrations may be below such values. On the other hand, NOx concentrations in the rural planetary boundary layer over the United States usually exceed 1 ppb. NOx concentrations of 5 to 10 ppb are typical of rural areas within more heavily populated areas in the United States and Europe. Empirical estimates based on O3 and NOx measurements at a site at 3 km elevation, that is, Niwot Ridge in Colorado, indicate that in summer 17 3 ppb of O3 is formed per 1 ppb of NOx when NOx concentrations are below 1 ppb. At lower elevation rural sites elsewhere in the United States, where NOx

874

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FIGURE 23.3 Frequency functions of duration of O3 concentrations in excess of 80 ppb over New Jersey and Connecticut for 2 years. From Rao (1988).

concentrations were within the 1 to 10 ppb range, 5 to 7 ppb of O3 was estimated to be formed per 1 ppb of NOx reacted. Since O3 is highly reactive with ground-level surfaces, it drops markedly in the evening. However, it can remain at elevated concentrations in the ambient air above the mixing layer. This elevated reservoir of O3 can then contribute to elevated ground level O3 on the following day as air mixing increases. This contributes to multiday summer episode exposures. Rao (1988) has shown that the likelihood of O3 >80 ppb continuing for 3 days or longer, once it has been in existence for 1 day, is high in the northeastern United States. This is illustrated in Fig. 23.3 for a typical year (1981) and a relatively high O3 exposure year (1983). The peak concentrations of O3 during a specific day at a specific location are determined largely by the baseline level in the air aloft, the photochemical production rate during the day, and the concentration of O3 scavenging chemicals such as nitric oxide (NO) and ethylene, and depends on the ambient ratio of reactive organic gases (ROGs) to NOx concentrations. When [ROG]/[NOx] is approximately 5–6, the two species have about an equal chance of reacting with hydroxy (OH) radical. If this ratio is much larger than 5 to 6, there is a shortage of NO that can be oxidized to NO2, and O3 production is controlled by the amount of NOx available. In this region, decreasing NOx leads to a decrease in the peak O3. However, when [ROG]/[NOx] is on the order of 5 or less, the ready availability of NOx makes O3 formation dependent on ROG. NO will scavenge O3 faster than it reacts with RO2, and NO2 will react with OH to form nitric acid. Decreasing NOx can lead to an increase in peak O3 as the efficiency of O3 formation increases. The daily formation of O3, in the absence of a substantial baseline level from the air aloft or upwind, leads to a relatively sharp daily peak in concentration, with a major part of the effective exposure taking place over a relatively few hours. However, in recent years, this type of exposure pattern has become relatively rare. In heavily populated regions, such as the eastern United States and western Europe, a typical daily plateau of exposure occurs after 10 a.m. and the maximum 8 h exposure is approximately 90% of the maximum 1 h exposure (Rombout et al., 1986). This type of exposure pattern is consistent with the hypothesis that relatively little of the exposure on a typical high exposure summer day is attributable to local

BACKGROUND ON EXPOSURES AND HEALTH-RELATED EFFECTS

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FIGURE 23.4 Three-day sequence of hourly O3 concentrations at Montague, MA SURE station showing locally generated midday peaks and long-range transport late peaks. From U.S. EPA (1986).

sources or amenable to local source control. Rather, the local generation of O3 represents a bump on a broad daily hump arising from a series of upwind sources and photochemistry. The size of the bump depends on the concentration of precursor reactants in the incoming air and the local increments of reactants. The broad humps can be attributed to the sum of the contributions of stratospheric O3 injections and O3 formed upwind and retained aloft for one or many days. The nature of contemporary O3 exposure is illustrated for a rural area of western Massachusetts in Fig. 23.4, showing both locally generated peaks and late-afternoon peaks from upwind population centers superimposed on a broad daily plateau (Lioy and Dyba, 1989). It clearly illustrates that the O3 exposure problem affects broad areas of the country and is not an urban problem only. 23.2.2

Ozone Exposures and Dosimetry

For O3, the only significant exposure route is inhalation, and exposure can be defined as the concentration at the nose and mouth. There have been few personal measurements of O3, and it is generally assumed that the concentrations that we breathe are the same as those measured at central monitoring sites. However, this assumption has limited validity. For time spent outdoors, local concentrations are reduced in the vicinity of heavy vehicular traffic due to scavenging by NO. However, less traffic areas downwind of the monitor may have a higher O3 concentration because of the enrichment of the air mass with motor vehicle exhaust precursor chemicals and active photochemistry. Thus, outdoor O3 concentrations can be either higher or lower than those measured at monitoring sites. Indoor concentrations of O3 are almost always substantially lower than those outdoors because of efficient scavenging by indoor surfaces and the lack of indoor sources. The only common indoor sources are copying machines and electrostatic air cleaners. Since most people spend more than 80% of their time indoors, their O3 exposure is much lower than estimates based on outdoor concentrations. Ozone exposure is only one determinant of O3 dose. Dose is also determined by the volumes of air inhaled and by the pattern of uptake of O3 molecules along the respiratory tract. When people work or exercise outdoors and increase their rate of ventilation, the

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contribution of outdoor exposure to total dose of O3 becomes the major determinant of total O3 dose. The dose to target tissues in the respiratory acini (the region from the terminal bronchioles through the alveolar ducts) increases even more with exercise than does total respiratory tract dose, since O3 penetration to distal lung airways increases with tidal volume and flow rate. Gerrity et al. (1988) measured the efficiency of O3 removal from inspired air by the extrathoracic and intrathoracic airways in healthy, nonsmoking young male volunteers for O3 concentrations of 100, 200, and 400 ppb for nose only, mouth only, and oronasal breathing, respectively, at 12 and 24 breaths/min. The mean extrathoracic removal efficiency for all measurements was 39.6 0.7%, and the mean intrathoracic removal efficiency was 91.0 0.5%. The effects of concentration, breathing frequency, and mode of breathing on removal efficiency while significant were relatively small. Gerrity et al. (1995) used a bronchoscope to sample air at various lung depths in healthy nonsmokers, with the distal end sequentially positioned at the bronchus intermedius (BI), main carina (CAR), upper trachea, and above the vocal cords. O3 concentration was measured continuously at each site using a rapid-responding O3 analyzer. The subjects breathed through a mouthpiece at 12 breaths/min. Integration of the product of the flow and O3 concentrations during inspiration and expiration provided the O3 mass passing each anatomic location during each phase of respiration. On inspiration, the fractional uptake of O3 by structures between the mouth and each location were 0.18 0.04 (SE), 0.27 0.02, 0.36 0.03, and 0.33 0.03 for above the vocal cords, upper trachea, CAR, and BI, respectively. A significant effect of location on uptake was found by an analysis of variance. Studies of the O3 uptake within the human respiratory tract have been conducted by Rigas et al. (2000) in tidal breathing. Male and female adults inhaled 200 or 400 ppb O3 while exercising at 20 L/min for 60 min or 40 L/min for 30 min. Fractional absorption ranged from 0.56 to 0.98, with an intersubject variability of approximately 10%. In the same laboratory, Asplund et al. (1996) used a continuous O3 exposure followed by an O3 bolus, and Rigas et al. (1997) used O3 boli following continuous NO2 and SO2 exposures. With continuous O3 exposure, the absorbed fraction of the bolus decreased, suggesting that biochemical substances on the airways were being depleted, whereas with continuous NO2 and SO2 exposures, the absorbed fraction of the O3 bolus increased, suggesting that the NO2 and SO2 exposures were increasing the availability of the biochemical substances that absorb O3. The tissues within the respiratory acini of humans, rabbits, guinea pigs, and rats receive the highest local dose from inhaled O3 according to the models developed by Miller and colleagues (Hatch et al., 1989; Miller et al., 1978b; Overton and Miller, 1987), with the dose in humans being about twice that in rats for the same exposure (Gerrity and Wiester, 1987), and with children having somewhat greater doses than adult humans (Overton and Graham, 1989). This comparative dosimetry is consistent with the greater effects of O3 on lung function seen for comparable exposures in humans than in rats (Costa et al., 1989). 23.2.3

Populations of Concern for Health Effects

In general, the NAAQSs have been established to protect against adverse health effects in the most sensitive subpopulation that is identifiable (Lippmann, 1987). For example, cardiovascular patients were of paramount concern in establishing the NAAQS for carbon monoxide (CO), which binds to hemoglobin and further reduces their already limited capacity to oxygenate the blood. Asthmatics were of special concern in establishing the sulfur dioxide (SO2) NAAQS because the concentrations required to produce comparable

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levels of bronchoconstriction were about an order of magnitude lower than those for normal people and because of the potential for a disabling or fatal bronchospasm being initiated by a transient high concentration of SO2. In the case of O3, no special functional responsiveness has yet been clearly demonstrated among the potentially more sensitive groups with preexisting disease (Lippmann, 1989a, 1989b, 1993). Thus, consideration is being given to healthy people who exercise regularly outdoors as a primary population of concern on the basis of their higher O3 exposures and doses. The EPA ozone staff papers have also identified people with asthma as a population of concern on the basis of reports of symptomatic responses, increased rates of visits to clinics, and hospital admissions at very low ambient O3 concentrations. 23.2.4

Health-Related Responses of Concern

As has been previously noted, O3 in ambient air has been associated with a variety of transient effects on the respiratory airways. Among the best documented of these changes are dose-related decrements in indices of forced expiratory flow capacity, which is reproducible in individuals and highly variable among the population. Increased rates of symptoms, clinic visits, and hospital admissions are other responses of concern with respect to peak exposures. More persistent physiological decrements associated with structural alterations of lung airways could also be considered adverse effects if they occurred in humans as a result of repetitive O3 exposures. Although human evidence is currently lacking, such effects have been produced in laboratory animals following chronic exposures. Thus, these effects are also of concern for human populations with high levels of chronic exposure.

23.3 EFFECTS OF SHORT-TERM EXPOSURES TO OZONE IN HUMANS 23.3.1

Respiratory Mechanical Function Responses

23.3.1.1 One- and Two-Hour Chamber Exposure Studies There are more data on respiratory function responses than on any other coincident responses to short-term O3 inhalation. Such functional responses can be obtained with noninvasive, readily performed protocols and can be detected at levels of exposure as low as or lower than any of the other well-established assays. The major debate about very small but statistically significant decrements in function from such studies is how to interpret their health significance (Lippmann, 1988). It is well established that the inhalation of O3 causes concentration-dependent mean decrements in exhaled volumes and flow rates during forced expiratory maneuvers, and that the mean decrements increase with increasing depth of breathing (Hazucha, 1987). There is a wide range of reproducible responsiveness among healthy subjects (Frampton et al., 1997; McDonnell et al., 1985a; Weinmann et al.,1995). Functional responsiveness to O3 is not greater, and usually lower, among cigarette smokers (Frampton et al., 1997; Kagawa, 1984; Shephard et al., 1983), older adults (Drechsler-Parks et al., 1987; McDonnell et al., 1993, 1995; Reisenauer et al., 1988), asthmatics (Koenig et al., 1987; Linn et al., 1980), and patients with chronic obstructive pulmonary disease (COPD) (Linn et al., 1983; Solic et al., 1982). An exception is that patients with allergic rhinitis had greater changes in airway resistance (McDonnell et al., 1987).

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A prospective confirmation of reduced responsiveness to O3 among asymptomatic cigarette smokers was produced by Emmons and Foster (1991). They measured respiratory function before and after 2 h of O3 at 400 ppb with light exercise in smokers before they stopped smoking and again after 6 months of not smoking. None was responsive to O3 exposure before smoking cessation. During smoking cessation, their mean baseline FEF25–75 was raised from 3.0 to 4.1 L/sec. For the subjects re-exposed to O3 6 months later, the exposure reduced their mean FEF25–75 from 3.9 to 3.0 L/sec. The subjects with the greatest improvement in FEF25–75 after withdrawal had the largest acute decrements after O3 exposure. Smoking cessation did not significantly increase FVC or forced expiratory volume in 1 s (FEV1), and O3 exposure after smoking cessation did not produce significant decrements in these respiratory function parameters. Weinmann et al. (1995) showed that O3-induced changes in FEF25–75 were unexplained and followed a different time-course than O3-induced changes in FVC. Their analysis indicated that intrinsic narrowing of the small airways might be a significant indicator of the functional response. While the results of some laboratory studies have indicated that responses in young females was greater than those in young males (Messineo and Adams, 1990), the largest study of both males and females did not find gender-related differences in responsiveness to O3 among either black or white adults (Seal et al., 1993). The first indications that the effects of O3 on respiratory function accumulate over more than 1 h were the observations of McDonnell et al. (1983) and Kulle et al. (1985) in chamber exposures to O3 in purified air for 2 h with the volunteers engaged in vigorous intermittent exercise. Significant function decrements observed after 2 h of exposure were not present at measurements made after 1 h. 23.3.1.2 Field Studies Spektor et al. (1998a) noted that children at summer camps with active outdoor recreation programs had greater decrements in lung function than children exposed to O3 at comparable concentrations in chambers for 1 or 2 h. Furthermore, their activity levels, although not measured, were known to be considerably lower than those of the children exposed in the chamber studies while performing very vigorous exercise. Since it is well established that functional responses to O3 increase with levels of physical activity and ventilation (Hazucha, 1987), the greater responses in the camp children had to be caused by other factors, such as greater cumulative exposure, or by the potentiation of the response to O3 by other pollutants in the ambient air. Cumulative daily exposures to O3 were generally greater for the camp children, since they were exposed all day long rather than for a 1 or 2 h period preceded and followed by exposure to clean air. Similar considerations apply to the studies of Kinney et al. (1988) and Hoek et al. (1993) of school children. In the Kinney et al. (1988) study in Kingston and Harriman, TN, lung function was measured in school on six occasions during a 2-month period in the late winter and early spring. Child-specific regressions of function versus maximum 1 h O3 during the previous day indicated significant associations between O3 and function, with coefficients similar to those seen in the summer camp studies of Lippmann et al. (1983), Spektor et al. (1988a, 1991, Higgins et al. (1990), and Hoek et al. (1993). Since children in school may be expected to have relatively low activity levels, the relatively high response coefficients may be related to potentiation by other pollutants or to a low level of seasonal adaptation. Kingston–Harriman is notable for its relatively high levels of aerosol acidity. As shown by Spengler et al. (1989), Kingston–Harriman has higher annual average and higher peak acid aerosol concentrations than other cities studied, that is, Steubenville, OH; St. Louis,


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MO; and Portage, WI. Alternatively, the relatively high response coefficients could have been caused by the fact that the measurements were made in the late winter and early spring. Linn et al. (1988) have shown evidence for a seasonal adaptation, and children studied during the summer may not be as responsive as children measured earlier in the year. This study will be discussed further Section 23.8. In a study of children with moderate to severe asthma at a summer camp in the Connecticut River Valley (Thurston et al., (2001)), the association between decrements in peak expiratory flow rates associated with ambient O3 concentrations were similar in magnitude to those reported by the same group of investigators for healthy children at other summer camps in northeastern United States (Spektor et al., (1988a, 1991). However, the level of physical activity of the asthmatic children, and hence their O3 intake, was much lower. Also, the asthmatic children have less reserve functional capacity. Thus, the level of health concern for such comparable functional decrements is much greater. Other recent studies of the effects of O3 on lung function in children in natural settings have also demonstrated O3-related functional decrements. Braun-Fahrlander et al. (1994) showed O3-related reductions in peak expiratory flow rate (PEFR) among 9–11-year-old Swiss children following 10 min of heavy exercise at peak O3 concentrations below 80 ppb. Neas et al. (1995) demonstrated O3-related reductions in PEFR between morning and evening in fourth- and fifth-grade children in Uniontown, PA, in relation to 12 h av. O3 below 88 ppb. Castillejos et al. (1995) studied the change in lung function following exercise out of doors for 7 1/2–11-year-old children in Mexico City who were repeatedly exposed to high ambient levels of O3 and particulate matter (PM). They had O3-related decrements in FVC, FEV1.0, FEF25–75, and FEV1.0/FVC when peak 1 h O3 exceeded 150 ppb. Field studies of functional responses of adults engaged in recreational activities outdoors in the presence of varying levels of O3 have also been performed. Spektor et al. (1988b) made pre- and postexercise respiratory function measurements on young adults who were engaged in daily outdoor exercise for about one-half hour per day in an area with regional summer haze but no local point sources. The magnitudes of the functional decrements per unit of ambient O3 concentration were similar to those observed in volunteers exposed while exercising vigorously for 1 or 2 h in controlled chamber exposure studies. Functional decrements in proportion to relatively low ambient O3 concentrations have also been reported for joggers in Houston, TX (Selwyn et al., 1985), competitive cyclists in The Netherlands (Brunekreef et al., 1994), hikers on Mount Washington in NH (Korrick et al., 1998), and agricultural workers in British Columbia (Brauer et al., 1996). 23.3.1.3 Prolonged Daily Exposures in Chambers The observations from the field studies in children’s camps stimulated Folinsbee et al. (1988) at the EPA Clinical Studies Laboratory in Chapel Hill, NC, to undertake a chamber exposure study of 10 adult male volunteers involving 6.6 h of O3 exposure at 120 ppb. Moderate exercise was performed for 50 min/h for 3 h in the morning and again in the afternoon. They found that the functional decrements become progressively greater after each hour of exposure, reaching average values of approximately 400 mL for forced vital capacity and approximately 540 mL for forced expiratory volume in 1 s by the end of the day. The effects were transient in the sense that there were no residual functional decrements on the following day. The decrements in FEV1 after 6.6 h of exposure at 120 ppb averaged 13.6% and were comparable to those seen previously in the same laboratory on similar subjects following 2 h of intermittent heavier exercise (68 L inhaled per minute for a total exercise time of 60 min) at an interpolated concentration of approximately 120 ppb. Assuming that the rate of ventilation was 10 L/min

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FIGURE 23.5 Mean FEV1 after each 50 min of exercise during exposures to O3 at 0 (open circles), 80 ppb (squares), 100 ppb (triangles), and 120 ppb (solid circles). Asterisks indicate significant reduction in FEV1 from corresponding values at 0 ppb. From Horstman et al. (1990).

between exercise periods, the total amount of O3 inhaled during 2 h of intermittent heavy exercise at 220 ppb (430 mg/m3) would be (60 min 0.068 m3/min þ 60 min 0.010 m3/ min) 430 mg/m3 ¼ 2.01 mg O3. The corresponding amount of O3 inhaled during 6.6 h of intermittent moderate exercise at 120 ppb would be (300 min 0.040 m3/min þ 100 min 0.010 m3/min) 235 mg/m3 ¼ 3.06 mg O3. Thus, the effect accumulated with time, but there was a temporal decay of the effect going on at the same time. Follow-up studies (in the same laboratory) by Horstman et al. (1990) were done on 21 adult males with 6.6 h exposures at 80, 100, and 120 ppb. The exposures at 120 ppb produced very similar responses, for example, a mean FEV1 decline of 12.3%, whereas those at 80 and 100 ppb showed lesser changes that also became progressively greater after each hour of exposure (Fig. 23.5). A further follow-up study using the same exposure protocol on 38 additional healthy young men was done by McDonnell et al. (1991) at 80 ppb. There was a mean FEV1 decline of 8.4%, which was similar to that seen by Horstman et al. (1990) at that concentration. The timescale for an effective O3 dose in relation to functional response was explored further by Hazucha et al. (1992) in exposures of healthy young adults lasting 8 h, with 30 min of exercise (@ 40 L/min) at the beginning of each hour. The O3 concentration rose from 0 to 240 ppb over the first 4 h and dropped back to zero over the second 4 h. The functional responses were compared with both sham exposures and constant 120 ppb exposures in the same subjects. By 4 h, the FEV1 changes from both O3 exposures were similar, and the largest decrement in FEV1, which occurred after 6 h of exposure, was about twice as large as that after 5 to 8 h of constant exposure at 120 ppb. The peak response faded by the end of 8 h and was not significantly greater than that produced by the constant 120 ppb exposure at the eighth hour. Another study looking into the integral effects of temporally varying exposures with the same integral exposure was performed by McKittrick and Adams (1995). Aerobically trained young adult men were exposed while exercising at 60 L/min to either 1 h @ 300 ppb O3 followed by 1 h of clean air; intermittent 1/2 h at 300 ppb and 0 ppb, or intermittent quarter hours at 300 and 0 ppb. The FEV1 decrements at the end of exposure to O3 were essentially the same, that is, 17.6, 17.0 and 17.9%. Larsen et al. (1991) modeled the data of Horstman et al. (1990) using multiple linear regressions on the mean responses at each hour for all three concentrations, but


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FIGURE 23.6 Comparison of mean O3-induced FEV1 decrements due to 6.6 hr exposures with mild exercise in the various studies cited in the inset.

excluding those with FEV1 decreases of less than 0.5%. With O3 concentration and duration of exposure as the only two independent variables, the model explained 95% of the variance of the dependent variable Z, a Gaussian transform of the percentage decrease in FEV1. In this model, the exponent of the exposure duration is 0.754. This further demonstrates that exposure time is almost equally important to exposure concentration in cumulative response when concentrations are in the range of normal peak ambient levels. Further evidence of the time scale for the biological integration of O3 exposure can be deduced from the rate at which the effects dissipate. In a study by Folinsbee and Hazuch (), young adult females were exposed to 350 ppb O3 for 70 min, including two 30 min periods of treadmill exercise at 40 L/min. Their mean decrement in FEV1 at the end of the exposure was 21%. After 18 h, their mean decrement was 4%, whereas at 42 h it was 2%. The large interindividual variability of O3-induced functional responses that is illustrated in Fig. 23.6 is not yet understood, and functional responses in individuals do not correlate well with the other responses that will be discussed below. Using the large EPA database, McDonnell et al. (1993) found that O3 concentration explained 31% of the variance in FEV1 responses, and subject age explained another 4%. The modeled influence of age is illustrated in Fig. 23.7. Upon further modeling of this large data set, McDonnell et al. (1997) reported that a sigmoid-shaped model was consistent with previous observations of O3 exposure–response (ER) characteristics and accurately predicted the mean response with independent data. Neither did they find that response was more sensitive to changes in C than in VE nor did they find convincing evidence of an effect of body size upon response, but response to O3 decreased with age. Using the data collected for 68 individuals exposed two or more times for 6.6 h, McDonnell et al. (1995) found that 47% of those exposed to 120 ppb had an FEV1 decrement of 10% or more. These analyses helped demonstrate that the respiratory function effects can accumulate over many hours and that an appropriate averaging time for transient functional decrements caused by O3 is 6 h. This was a major factor for the change in the

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FIGURE 23.7 Predicted mean decrements in forced expiratory volume in L(DFEV1) following 2 h esposures to ozone while undergoing heavy intermittent exercise for three ages. (Note: to convert DFEV1 to % DFEV1, multiply by 22.2%.). From McDonnell et al. (1993).

averaging time for the primary O3 NAAQS from 1 h to 8 h. Another factor was the recognition that O3 exposures in ambient air can have broad peaks with 8 h averages equal to approximately 90% of the peak 1 h averages (Rombout et al., 1986). 23.3.2

Effects on Athletic Performance

It has been four decades since epidemiological evidence suggested that the percentage of high school track team members failing to improve performance increased with increasing oxidant concentrations the hour before a race (Wayne et al., 1967). The effects may have been related to increased airway resistance or to associated discomfort, which may have limited motivation to run at maximal levels. Controlled exposure studies of heavily exercising competitive runners have demonstrated decreased function at 200 to 300 ppb (Adams and Schelegle, 1983; Savin and Adams, 1979). At 210 ppb O3, Folinsbee et al. (1984) reported symptoms as well in seven distance cyclists exercising heavily (VE ¼ 81 L/min). Some studies have shown reduced performance at lower O3 concentrations. Schelegle and Adams (1986) exposed young male adult endurance athletes to 120, 180, and 240 ppb O3 whereas the exercised at a mean VE of 54 L/min for 30 min, followed by a mean VE of 120 L/min for an additional 30 min. Although they all completed the protocol for filtered air (FA) exposure, some of them could not complete it for the 120, 180, and 240 ppb exposures. Linder et al. (1988) also found that maximum performance time was reduced for their 16–28 min progressive maximum exercise for VE of 30–120 L/min in young adults when O3 was present. For example, performance was reduced 11% in females exposed to 130 ppb O3. 23.3.3

Symptomatic Responses

Respiratory symptoms have been closely associated with group mean pulmonary function changes in adults acutely exposed in controlled exposures to O3 and in ambient air containing O3 as the predominant pollutant. However, Hayes et al. (1987) found only a weak-tomoderate correlation between FEV1 changes and symptoms severity when the analysis was conducted using individual data. In controlled 2 hr O3 exposures, McDonnell et al. (1983) reported that some heavily exercising adult subjects experienced cough, shortness of breath, and pain on deep


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inspiration at 120 ppb O3, although the group mean response was statistically significant for cough only. Above 120 ppb O3, respiratory and nonrespiratory symptoms include throat dryness, chest tightness, substernal pain, cough, wheeze, pain on deep inspiration, shortness of breath, dyspnea, lassitude, malaise, headache, and nausea. The prolonged exposure studies involving 6.6 h of exposure at concentrations between 80 and 120 ppb also produced significant increases in respiratory symptoms including cough and pain on deep inspiration (Koenig et al., 1987; Linn et al., 1980). Linder et al. (1988) reported that brief exposures (16–28 min) to 120 to 130 ppb O3 at high ventilatory rates (30–120 L/min) produced symptoms of irritation and cough in young adults. Although O3 causes symptomatic responses in adults at current peak levels, such responses do not occur in healthy children (Avol et al., 1985, 1987). Children (ages 8– 11) exposed for 2.5 h at 120 ppb O3 while intermittently exercising (VE ¼ 39 L/min) showed small but statistically significant decreases in FEV1 but showed no changes in frequency or severity of cough compared to controls (McDonnell et al., 1985a, 1985b). Similarly, adolescents (age 12–15) continuously exercising (VE ¼ 31–33 L/min) during exposure to 144 ppb mean O3 in ambient air showed no changes in symptoms despite statistically significant decrements in group mean FEV1 (4%), which persisted at least 1 h during postexposure resting (Avol et al., 1985). These laboratory results are consistent with the results obtained in a series of field studies of healthy children at summer camps, which failed to find any symptomatic responses despite the occurrence of relatively large decrements in function that were proportional to the ambient O3 concentrations (Spektor et al., 1988a). In a study by Hoek and Brunekreef (1995) of a general population sample of 300 children aged 7–11 years who had shown functional responses to O3 in ambient air, there were no responses in terms of symptoms based on diaries maintained by their parents. In panels of 300 healthy children in the Harvard six-cities study, diaries of respiratory symptoms were kept over a 1-year period. In single pollutant models for the April–August period, there was a significant association between O3 and the incidence of cough that was independent of other measured pollutants (Schwartz et al., 1994). For a group of 7–9-year-old children in Mexico City who were repeatedly exposed to high concentration of O3 and PM, Castillejos et al. (1995) reported that mean O3 in the previous 48 h was associated with a child’s report of cough or phlegm, while mean O3 in the previous day or week was not. For a panel of 71 asthmatic, 5–7-year-old children in Mexico City. respiratory symptoms (coughing, phlegm production, wheezing, and difficult breathing) and the frequency of lower respiratory illness on the same day were associated with both O3 and PM10 (Romieu et al., 1996). In a study of children with moderate-to-severe asthma in the Connecticut River Valley, where O3 exposures were much lower than those in Mexico city, Thurston et al. (1997) fount that respiratory symptoms were significantly associated with O3. Other epidemiology studies have provided evidence of qualitative associations between ambient oxidant levels >0.10 ppm and symptoms in children and young adults, such as throat irritation, chest discomfort, cough, and headache (Hammer et al., 1974; Makino and Mizoguchi, 1975). Thus, symptoms reported in individuals exposed to O3 in purified air are similar to those found in individuals exposed to ambient air except for eye irritation, a common symptom associated with exposure to photochemical oxidants, which has not been reported for controlled exposures to O3 alone. Other oxidants, such as aldehydes and peroxyacetyl nitrate (PAN), are primarily responsible for eye irritation and are generally found in atmospheres containing higher ambient O3 levels (Altshuller, 1977; National Research Council, 1977).

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There have been several studies reporting associations between ambient photochemical oxidant pollution and exacerbation of asthma (Holguin et al., 1985; Schoettlin and Landau, 1961; Whittemore and Korn, 1980), but the role of specifically O3 and the nature of the exposure–response relationships remain poorly defined. Respiratory symptoms in healthy young adult females (student nurses) in Los Angeles, in relation to ambient pollution levels, were monitored by Hammer et al. (1974). Schwartz and Zeger (1990) reexamined the original diaries from this study, which contained smoking and allergy histories as well as symptom reports that had never been analyzed. Diaries were compiled daily and collected weekly for as long as 3 years. Air pollution was measured at a monitoring location within 2.5 miles of the school. Incidence and duration of a system were modeled separately. Photochemical oxidants (74 ppb) were associated with increased risk of chest discomfort (odds ratio (OR) ¼ 1.17; p < 0.001) and eye irritation (OR ¼ 1.20; p < 0.001). Ostro et al. (1993) recorded the respiratory symptoms in nonsmoking adults residing in Southern California. Participants recorded the daily incidence of several respiratory symptoms over a 6 month period between 1978 and 1979. Ambient concentrations of O3, SO42 , and other air pollutants were measured. Using a logistic regression model, the authors found a significant association between the incidence of lower respiratory tract symptoms and 7 h O3 (OR ¼ 1.32; 95% confidence interval (CI): 1.14–1.52, for a 100 ppb change), and SO42 (OR ¼ 1.30; 95% CI: 1.09–1.54, for a 10 mg/m3 change), but no association was found with coefficient of haze, a more general measure of PM. The existence of a gas stove in the home was also associated with lower respiratory tract symptoms (OR ¼ 1.23; 95% CI: 1.03–1.47). The effects of O3 were greater in the subpopulation without a residential air conditioner. In addition, O3 had a greater effect on individuals with a preexisting respiratory infection. Desqueyroux et al. (2002) studied symptomatic responses to community air pollutants among patients with COPD. During a 14 month period, Parisian adults with severe COPD were monitored by their physicians. Daily levels of four air pollutants were provided by an urban air quality network. Exacerbation of COPD was associated only with O3 (OR ¼ 1.44 for a 5 ppb increase in O3; 95% CI: 1.14, 1.82), with a lag of 2–3 days. The effect of O3 was greater in patients whose CO2 pressure (PaCO2) was higher than 43 mmHg (OR ¼ 1.83; 95% CI: 1.36, 2.47). 23.3.3.1 Effects on Airway Reactivity Exposure to O3 can also alter the responsiveness of the airways to other bronchoconstrictive challenges as measured by changes in respiratory mechanics. For example, Folinsbee et al. (1988) reported that airway reactivity to the bronchoconstrictive drug methacholine for the group of subjects as a whole was approximately doubled following 6.6 h exposures to 120 ppb O3. Airway hyperresponsiveness (to histamine) had previously been demonstrated but only at O3 concentrations 400 ppb (Holtzman et al., 1979; Seltzer et al., 1986). On an individual basis, Folinsbee et al. (1988) found no apparent relationship between the O3-associated changes in methacholine reactivity and those in FVC or FEV1. On the other hand, Aris et al. (1991) reported a closer relationship, more similar to reported responses to inhaled H2SO4 aerosol, where changes in function correlated closely with changes in reactivity to carbachol aerosol, a bronchonconstrictive drug (Utell et al., 1983). The O3-associated changes in bronchial reactivity may predispose individuals to bronchospasm from other environmental agents such as acid aerosol and naturally occurring aeroallergens.


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Tests by Horstman et al. (1990), involving 6.6 h exposures to 80, 100, and 120 ppb, produced 56%, 89%, and 121% increases in methacholine responsiveness, respectively. Increased responsiveness to methacholine was also seen in the Folinsbee and Hazuch (1989) study with 1 h O3 exposure at 350 ppb. An increased responsiveness to histamine was seen by Gong et al. (1988) in one of 17 competitive cyclists exposed at 120 ppb for 1 h at VE of 89 L/ min followed by 3–4 min at 150 L/min. At 200 ppb, responsiveness increased in 9 of 17 subjects. McDonnell et al. (1987) found increased histamine responsiveness in 26 young adult males with allergic rhinitis after O3 at 180 ppb during 2 h of exercise at 64 L/min. Jorres et al. (1996) exposed 24 subjects with mild stable allergic asthma, 12 subjects with allergic rhinitis without asthma, and 10 healthy subjects to 250 ppb O3 or FA for 3 h with intermittent exercise. They determined the concentration of methacholine (PC20FEV1) and the dose of allergen (PD20FEV1) producing a 20% fall in FEV1. In subjects with asthma, FEV1 decreased by 12.5 2.2%, PC20FEV1 of methacholine by 0.91 0.19 doubling concentrations and PD20FEV1 of allergen by 1.74 0.25 doubling doses after O3 compared with sham exposure to FA. The changes in lung function, methacholine, and allergen responsiveness did not correlate with each other. In subjects with rhinitis, mean FEV1 decreased by 7.8% and 1.3% when O3 or FA, respectively, were followed by allergen inhalation. 23.3.3.2 Effects on Airway Permeability Kehrl et al. (1987) studied the effects of inhaled O3 on respiratory epithelial permeability in healthy, nonsmoking young men. They were exposed for 2 h to purified air and 400 ppb ozone while performing intermittent treadmill exercise at 67 L/min. Specific airway resistance (SRaw) and FVC were measured before and at the end of exposures. Seventy-five minutes after the exposures, the pulmonary clearance of a radioisotope-labeled organic molecule, that is, [99m Tc]DTPA, was measured as an index of epithelial permeability. O3 exposure caused respiratory symptoms in all eight subjects and was associated with a 14 2.8% (mean SE) decrement in FVC (p < 0.001) and a 71 22% increase in SRaw (p ¼ 0.04). Compared to the air exposure day, seven of the eight subjects showed increased [99m Tc]DTPA clearance after the O3 exposure, with the mean value increasing from 0.59 0.08 to 1.75 0.43%/min (p ¼ 0.03). Thus, O3 exposure sufficient to produce decrements in the respiratory function of human subjects also causes an increase in permeability. An increased permeability could facilitate the uptake of other inhaled toxicants and/or the release of inflammatory cells such as neutrophils onto the airway surfaces. Foster and Stetkiewicz (1996) studied the influence of O3 on lung permeability in healthy subjects at 18–20 h after 2 h exposures at 150 and 350 ppb. Permeability was measured in terms of the clearance rate of a water-soluble aerosol containing 99m Tc-labeled DTPA (diethylamine pentaacetic acid). Based on a sequence of g-camera measurements of 99m Tc clearance from the lungs, they concluded that 99m Tc-DTPA clearance from the lung periphery and apexes was significantly increased by O3 but changes in clearance for the base of the lung were not significant. The FEV1 at the late time after O3 was slightly but significantly reduced ( 2.1%) from pre-exposure levels. There was no relationship between the functional changes observed acutely after exposure to O3 and subsequent changes in 99m Tc-DTPA clearance or FEV1 observed at the late period. These results suggest that epithelial permeability of the lung is altered 18–20 h post-O3; this injury is regional, and the lung base appears to have a different time course of response or is in an adapted state with respect to O3 exposure. 23.3.3.3 Effects on Airway Inflammation Seltzer et al. (1986) showed that O3-induced airway reactivity to methacholine is associated with neutrophil influx into the airways and

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with changes in cyclooxygenase metabolites of arachidonic acid. For 2 h exposures to O3 at 400 ppb with intermittent exercise, the bronchoalveolar lavage (BAL) fluid had increased prostaglandins E2 and F2a and thromboxane B2 3 h after the O3 exposure. Reports of Koren et al. (1989) and Devlin et al. (1991) also described inflammatory and biochemical changes in the airways following O3 exposure. In the initial studies, subjects were exposed to 400 ppb for 2 h while performing intermittent exercise at a ventilation of 70 L/min to examine cellular and biochemical responses in the airways. The BAL was performed 18 h after the O3 exposure. An 8.2-fold increase in polymorphonuclear leukocytes (PMNs or neutrophils) was observed after ozone exposure, confirming the observations of Seltzer et al. (1986). Twofold increases in protein, albumin, and IgG were indicative of increased epithelial permeability, as previously suggested by the [99m Tc]DTPA clearance studies of Kehrl et al. (1987). In addition to confirmation of the Kehrl et al. (1987) findings, Koren et al. (1989) provided evidence of stimulation of fibrogenic processes including increases in fibronectin (6.4), tissue factor (2.1), factor VII (1.8), and urokinase plasminogen activator (3.6). There was a twofold increase in the level of prostaglandin E2 and a similar elevation of the complement component C3a. Levels of leukotrienes C4 and B4 were not affected. Devlin et al. (1991) reported that a significant inflammatory response, as indicated by increased levels of PMN, was also observed in BAL fluid from subjects exposed to either 80 or 100 ppb O3 for 6.6 h. As illustrated in Fig. 23.8, the 6.6 h at 100 ppb O3 produced a 3.8-fold increase in PMNs at 18 h after the exposure, whereas the 6.6 h at 80 ppb produced a 2.1-fold increase. The amounts of O3 inhaled in the 80 and 100 ppb protocols were approximately 2.0 and 2.5 mg and approximately 3.6 mg in the 400 ppb protocol. Thus, the effect of concentration was apparently somewhat greater than that of exposure duration. The significant increase in PMNs at a concentration as low as 80 ppb suggests that lung inflammation from inhaled O3 has no threshold down to ambient background O3 levels. The inflammatory process caused by O3 exposure is promptly initiated (Seltzer et al., 1986) and persists for at least 18 h (Koren et al., 1989). The time-course of this inflammatory response, and the O3 exposures necessary to initiate it, however, has not yet been fully

FIGURE 23.8 Range of subject response 18 h after 6.6 h of O3 exposure at 100 ppb (closed circles) or 80 ppb (open circles). Squares indicate the mean changes (SE). From Devlin et al. (1991).


887

elucidated. Furthermore, these studies demonstrate that cells and enzymes capable of causing damage to pulmonary tissues were increased, and the proteins that play a role in the fibrotic and fibrinolytic processes were elevated as a result of O3 exposure. Scannell et al. (1996) studied a group of asthmatic subjects exposed to O3 using the same exposure protocol previously used by the same investigators for 81 healthy subjects. They reported no significant differences in lung function responses and a trend toward higher airway resistance (p < 0.13). By contrast, the asthmatic subjects had significantly greater (p < 0.05) O3-induced increases in inflammatory end points (% neutrophils and total protein) in bronchoalveolar lavage fluid (BALF) as compared to 20 of the normal subjects who also underwent bronchoscopy. Prolonged inflammatory processes following repetitive exposures to O3 in ambient air were reported by Kinney et al. (1996) in terms of reduced release of reactive oxygen species, increased levels of LDH, IL-8, and PGE2 in the BAL. Interpretation of the nature and significance of the inflammatory responses following short-term O3 exposures is difficult without knowledge of the cumulative effects that may be triggered by repetitive episodes of lung inflammation. The relation of the inflammatory responses, if any, to the well-studied respiratory function responses also remains unknown. We do know that these responses are poorly correlated. Balmes et al. (1996) tested the hypothesis that changes in lung function induced by O3 are correlated with indices of respiratory tract/injury inflammation. They exposed healthy subjects, on separate days, to O3 (0.2 ppm) and filtered air for 4 h during exercise. Symptom questionnaires were administered before and after exposure, and pulmonary function tests (FEV1, FVC, and Sraw) were performed before, during, and immediately after each exposure. Fiberoptic bronchoscopy, with isolated left main bronchus proximal airway lavage (PAL) and bronchoalveolar lavage (bronchial fraction, the first 10 ml of fluid recovered) of the right middle lobe, was performed 18 h after each exposure. The PAL, bronchial fraction, and BAL fluids were analyzed for total and differential cell counts, total protein, fibronectin, interleukin-8 (IL-8), and granulocyte– macrophage colony-stimulating factor (GM-CSF) concentrations. The study population was divided into two groups, least sensitive (n ¼ 12; mean O3-induced change in FEV1 ¼ 7.0%) and most sensitive (n ¼ 8; mean O3-induced change in FEV1 ¼ 36.0%). They found a highly significant O3 effect on Sraw and lower respiratory symptoms for all subjects combined but no significant differences between the least and most sensitive groups. O3 exposure increased significantly percent neutrophils in PAL; percent neutrophils, total protein, and IL-8 in bronchial fraction (p < 0.001, p < 0.001, and p < 0.01, respectively); and percent neutrophils, total protein, fibronectin, and GM-CSF in BAL for all subjects combined; there were no significant differences, however, between least and most sensitive groups. Thus, levels of O3-induced symptoms and respiratory tract injury/inflammation were not correlated with the magnitude of decrements in FEV1 and FVC. A similar conclusion was drawn by Torres et al. (1997), who studied whether individuals who differed in lung function responsiveness to O3, or in smoking status, also differed in susceptibility to airway inflammation. Healthy subjects were selected on the basis of responsiveness to a classifying exposure to 220 ppb O3 for 4 h with exercise (responders, with a decrease in FEV1 > 15%; and nonresponders, with a decrease in FEV1 < 5%). Three groups were studied: nonsmoker-nonresponders (n ¼ 12), nonsmoker-responders (n ¼ 13), and smokers (n ¼ 13; 11 nonresponders and two responders). Each subject underwent two exposures to O3 and one to air, separated by at least 3 weeks; bronchoalveolar and nasal lavages were performed on three occasions: immediately (early) and 18 h (late) after O3 exposure, and either early or late after air exposure. Recovery of PMNs increased progressively in all groups, and by up to 6-fold late after O3 exposure. IL-6 and IL-8 increased early

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(by up to 10-fold and up to 2-fold, respectively), and correlated with the late increase in PMN. Lymphocytes, mast cells, and eosinophils also increased late after exposure. Thus, O3-induced airway inflammation was independent of smoking status or airway responsiveness to O3. Alexis et al. (2000) used indomethacin pretreatment prior to O3 exposure to investigate the role that cyclooxygenase (COX) metabolites of arachidonic acid might play. They reported that COX metabolites contribute to restrictive-type changes in normal and obstructive-type changes in small airways in asthmatic subjects. Grievink et al. (1999) reported that 100 mg of vitamin E and 500 mg of vitamin E provided partial protection against O3-related function decrements in adult Dutch cyclists with O3 concentrations up to 93 ppb during the exercise. For 9-year-old children with moderate-tosevere asthma in Mexico City, with 8 h average O3 concentrations up to 184 ppb, daily supplements of 50 mg of Vitamin E and 250 mg of Vitamin C modulated the pulmonary effects of O3 (Romieu et al., 2002). However, 2 weeks of pretreatment of inhaled budesonide (a corticosteroid) with 800 mg twice a day provided no protection against inhaled O3 in terms of either pulmonary function, methacholine reactivity, or neutrophil recruitment (Nightingale et al., 2000). Holz et al. (1999) reported that respiratory function and O3-induced airway inflammatory changes differed between individuals, both for healthy and asthmatic subjects, were reproducible but were not related to each other. Vagaggini et al. (2002) studied subjects with mild asthma, as indicated by a methacholine challenge. They exposed them to an allergen 24 h before inhaling O3. The O3 exposure increased the percentage of eosinophils but not of PMNs in induced sputum above that associated with the allergen challenge alone. Samet et al. (2001) studied the pulmonary effects of O3 on healthy adults with and without dietary supplementation of antioxidants and found that the antioxidants reduced the O3induced functional decrements but not its effect on increasing PMNs and IL-6 in lavage fluid. Inflammatory reactions occur in the nasal passages as well as in the lungs. Graham et al. (1988) exposed 41 subjects to either filtered air or 500 ppb O3 for 4 h for 2 consecutive days. Nasal lavages (NLs) were taken before and immediately after each exposure and 22 h after the last exposure. Lavage PMN counts increased significantly (p ¼ 0.005) in the O3-exposed group, with 3.5-, 6.5-, and 3.9-fold increases over the air-exposed group at the post-1, pre-2, and post-2 time points, respectively. Graham and Koren (1990) compared the cellular changes detected in NL with those detected in BAL taken from the same individual. Subjects were exposed to either filtered air or 400 ppb O3, with exercise, for 2 h. The NL was done prior to, immediately after, and 18 h postexposure; the BAL was done only at 18 h postexposure. A significant increase in PMNs was detected in the NL immediately postexposure to O3 (7.7-fold increase; p ¼ 0.003) and remained elevated in the 8 h postO3 NL (6.1-fold increase; p < 0.001). A similar increase in PMNs was detected in the BAL 18 h after exposure to O3 (6.0-fold increase; p < 0.001). The albumin levels in the NL and BAL were also similarly increased 18 h after O3 (3.9-fold and 2.2-fold, respectively). Although a qualitative correlation in the mean number of PMNs existed between the upper and lower respiratory tracts after O3, a comparison of the NL and BAL PMNs from each individual showed a significant quantitative correlation for the air data (r ¼ 0.741; p ¼ 0.014) but not for the O3 data (r ¼ 0.408; p ¼ 0.243). The utility of this approach at low ambient levels of O3 was demonstrated by Frischer et al. (1993). They studied nasal airways inflammation after O3 exposure in children by repeated NL from May to October 1991. During this period, five to eight NLs were performed on each child. On day 14 following “high” O3 (>90 ppb), 148 NLs were performed and on day 10 following “low” O3 (10 mmHg increase), which were not significantly different between the two groups. These results suggest that O3 exposure can increase myocardial work and impair pulmonary gas exchange to a degree that might be clinically important in persons with significant pre-existing cardiovascular impairment, with or without concomitant lung disease. 23.4 FACTORS AFFECTING THE VARIABILITY OF RESPONSIVENESS IN HUMANS Although there is a great deal of knowledge about O3 exposure–respiratory function response in humans, as summarized above, we still know very little about the mechanisms responsible for the responses. Other irritants, such as SO2, NO2, and H2SO4, produce greater responses among asthmatics than among healthy human subjects but, as indicated previously, this is not true for O3. For other irritants, functional responses correlate with responsiveness to bronchoconstrictor challenge. For example, Utell et al. (1983) found a high correlation between reactivity to inhaled carbachol and responsiveness to inhaled H2SO4 in asthmatics (r ¼ 0.90, p< 0.001), whereas Horstman et al. (1986) reported that methacholine (MCh) reactivity and SO2 response were significantly but weakly correlated

FACTORS AFFECTING THE VARIABILITY OF RESPONSIVENESS IN HUMANS

891

(r ¼ 0.31). Although both functional decrements and bronchial responsiveness are produced by O3 exposure, Folinsbee et al. (1988) and Horstman et al. (1990) found no apparent relationship between these responses for individual subjects. On the other hand, Aris et al. (1991) screened healthy, nonsmoking volunteers for their functional responsiveness to 3 h of exposure to 200 ppb O3 at a ventilatory rate of 40 L/min and found that the MCh responsiveness of 10 O3-sensitive subjects (PC100 ¼ 3.0 0.8) was significantly greater than that of 10 O3-nonsensitive subjects (PC100 ¼ 18.7 4.5). Beckett et al. (1985) examined the effect of atropine, a muscarinic receptor blocker, on responses to exposure to 400 ppb O3. Atropine pretreatment prevented the significant increase in airway resistance with O3 exposure and partially blocked the decrease in forced expiratory flow rates, but it did not prevent a significant fall in FVC, changes in respiratory frequency, and tidal volume, or the frequency of reported respiratory symptoms. These results suggest that the increase in pulmonary resistance during O3 exposure is mediated by a parasympathetic mechanism and that changes in other measured variables are mediated, at least partially, by mechanisms not dependent on muscarinic cholinergic receptors of the parasympathetic nervous system. Gong et al. (1988) studied the contribution of b-adrenergic mechanisms to the acute airway responses to O3 in a study in which symptoms, pulmonary function, exercise performance, and postexposure histamine bronchoprovocation were studied in nonasthmatic athletes exposed to 210 ppb O3 during heavy continuous exercise, with mean minute ventilation (VE) 80 L/min for 60 min, followed by a maximal sprint (peak VE > 140 L/min) until exhaustion. Each subject was exposed randomly to either 210 ppb O3 or filtered air during the four single-blinded exposure sessions. Albuterol pretreatment resulted in modest but significant bronchodilation compared to placebo. However, albuterol did not prevent O3induced respiratory symptoms, decrements in FVC, FEV1, and maximum midexpiratory flow rate (FEF25–75), and positive histamine challenges compared to that with placebo and O3. There were statistically no significant differences in the metabolic data or ride times across all drugs and exposures, although the peak VE was significantly lower with O3 than FA regardless of drug. The results indicate that acute pretreatment with inhaled albuterol is unable to prevent or ameliorate O3-induced symptoms and alterations in pulmonary function and exercise performance. The contribution of b-adrenergic mechanisms in the acute airway responses to O3 appears to be minimal. In their study on bronchial hyperresponsiveness to O3 exposure, Seltzer et al. (1986) found significant increases in the concentration of prostaglandins E2 and F2a and thromboxane B2 in bronchoalveolar lavage fluid. Prostaglandins E2 and F2a stimulate pulmonary neural afferents that initiate several responses characteristic of acute O3 exposure (Coleridge et al., 1976; Roberts et al., 1985), suggesting that the release of prostaglandins in the lung may be involved in routinely observed pulmonary function decrements and perhaps in altered exercise ventilatory pattern and reported subjective symptomotology. Schelegle et al. (1987) studied whether O3-induced pulmonary function decrements could be inhibited by the prostaglandin synthetase inhibitor, indomethacin, in healthy human subjects. College-age males completed six 1 h exposure protocols with workloads set to elicit a VE of 60 L/min, with no drug, placebo, and indomethacin pretreatments, with filtered air and O3 (350 ppb) exposures within each pretreatment. Exposures consisted of 1 h exercise on a bicycle ergometer. Significant differences were found for comparisons of no drug versus indomethacin and of placebo versus indomethacin, suggesting that cyclooxygenase products of arachidonic acid, which are sensitive to indomethacin inhibition, play a

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prominent role in the development of pulmonary function decrements consequent to acute O3 exposure. In a similar study, Ying et al. (1990) administered indomethacin for 4 days to young adult male nonsmokers prior to 2 h O3 exposures with intermittent exercise at 400 ppb to determine if it would alter their O3 responsiveness as well as their lung function. For subjects who had detectable serum levels of indomethacin and significant responses to methacholine on the sham exposure day, the indomethacin attenuated the O3-induced decrements in lung function but did not attenuate the O3-induced responsiveness to methacholine. They concluded not only that the O3-induced decrements in respiratory function are mediated by cyclooxygenase products but also that the O3-induced increase in airway reactivity occurs by some other mechanism. The mechanism by which the release of cyclooxygenase products in the lung leads to pulmonary function decrements in humans upon O3 exposure remains undefined. Available data indicate that O3-induced pulmonary function decrements and ventilatory pattern changes are neurally mediated (Lee et al., 1979; Hazucha et al., 1989). Hazucha et al. (1989) concluded that O3 inhalation stimulates airway receptors, which leads to an involuntary inhibition of full inspiration, reduction in FVC, and a concomitant decrease in maximal expiratory flow rates in humans. The observation that cyclooxygenase products stimulate neural afferents in the lung (Coleridge et al., 1976; Roberts et al., 1985), combined with the observation of reduced O3-induced pulmonary function decrements after indomethacin pretreatment, suggests that cyclooxygenase products released consequent to O3induced tissue damage stimulate neural afferents in the lung, which results in the observed pulmonary function decrements.

23.5 STUDIES OF POPULATIONS EXPOSED TO OZONE IN AMBIENT AIR Observational studies of the influence of O3 on human health are often difficult to interpret because the population is also exposed to other pollutants in the ambient air that could affect the responses observed or to other environmental challenges that may produce comparable effects, such as environmental tobacco smoke, other pollutants in indoor air, and allergens found in indoor and outdoor air. For time-series studies of daily mortality and admissions to emergency departments, hospital admissions, and other health service providers, appropriate corrections need to be made for ambient temperature, which can covary with both O3 concentrations and health effect indices. 23.5.1

Mortality

Recent studies examining the possible influence of O3 on daily mortality have reported independent effects of O3 in multiple regression analyses. Sartor et al. (1995, 1997) found that O3 affected mortality during the summer of 1994 both for the individuals of all age groups and for the elderly in Belgium and that temperature potentiated the response to O3. Verhoeff et al. (1996) examined daily mortality in Amsterdam, The Netherlands, for the period of 1986–1992 and reported that O3 with a 2-day lag was positively associated with mortality, as were current day black smoke (BS) and PM10. There was no association with SO2 or CO. Anderson et al. (1996) studied air pollution and daily mortality in London, England, during 1987–1992. They reported that both same-day O3 and BS were independently associated with all causes of mortality, which was greater on warm days, and

STUDIES OF POPULATIONS EXPOSED TO OZONE IN AMBIENT AIR

893

independent of the effects of other pollutants. O3 was also significantly associated with mortality caused by cardiovascular and respiratory diseases. Touloumi et al. (1997) performed a combined analysis of daily mortality for six western and central European cities participating in the project Air Pollution and Health: A European Approach (APHEA). They reported that a 25 ppb increase in daily 1 h max. O3 concentration was associated with a 2.9% increase in the number of deaths, and the effect was independent of the BS concentration change and consistent across the cities. Thurston and Ito (2001) re-examined the data from a number of earlier time-series mortality studies that had not adequately corrected for ambient temperature. For all of the total mortality–air pollution time-series studies considered, the combined analysis yielded a relative risk (RR) ¼ 1.036 per 100 ppb increase in daily 1 h maximum O3 (95% CI: 1.023– 1.050). However, the subset of studies that specified the nonlinear nature of the temperature– mortality association yielded a combined estimate of RR ¼ 1.056 per 100 ppb (95% CI: 1.032–1.081) indicating that past time-series studies using linear temperature–mortality specifications underpredicted the mortality effects of O3 air pollution. For Detroit, MI, an illustrative analysis of daily total mortality during 1986–1990 also indicated that the model weather specification choice could influence the O3 health effect estimates. Results were intercompared for alternative weather specifications. Nonlinear specifications of temperature and relative humidity (RH) yielded lower intercorrelations with the O3 coefficient, and larger O3 RR estimates, than a base model employing a simple linear spline of hot and cold temperature. They concluded that, unlike for PM mass, the mortality effect estimates derived by time-series analyses for O3 can be sensitive to the way weather is addressed in the model. Generally, they found that the O3-mortality effect estimate increased in size and statistical significance when the nonlinearity and the humidity interaction of the temperature–health effect association were incorporated into the model weather specification. In recent years, many studies have been focused on the associations between short-term O3 exposures and daily mortality rates in urban centers. The National Mortality and Morbidity Air Pollution Study (NMMAPS) used EPA’s Atmospheric Information Retrieval System (AIRS) data on ambient O3 from 95 United States communities and publicly available daily mortality data in a preselected analytical model. As shown in Fig. 23.9, a positive association was found in all but 2 communities, and a statistically significant association was shown for 7 communities and for 95 communities as a whole (U.S. EPA, 2006). As shown in Fig. 23.10, the 95-community effect was strongest on the same day, and highly significant on 1- and 2-day lags, as well as being even stronger when the distributed lag over 6 days was considered. WHO-EURO commissioned a meta-analysis of time-series studies of the associations between ambient O3 and daily mortality in more than 80 studies published between 1996 and 2001. The results are summarized in Table 23.3. The relationship between acute effects of O3 on mortality was reinforced by the recent publication of four meta-analyses (Bell et al., 2005; Gryparis et al., 2005; Ito, 2005; Levy et al., 2005) that were coherent in showing a significant association between O3 and short-term mortality, which is not confounded by other pollutants (including particulate matter), temperature, weather, season and strategy of modeling. Increases in total mortality have been observed in a concentration as low as 75 mg/m3 (1 h mean) (Gryparis et al., 2005). While O3 does appear to have a significant impact on daily mortality rates, especially in the warmer months of the year, its average long-term concentration has not been found to have a significant influence on annual mortality rate. Pope et al. (2002) found fine particles to be associated with significant increase in cardiovascular and lung cancer mortality.

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FIGURE 23.9 Bayesian city-specific and national average estimates for the percent change (95% CI) in daily mortality per 10 ppb increase in 24 h average. O3 in the previous week using a constrained distributed lag model for 95 U.S. communities (NMMAPS), arranged by size of the effect estimate. Source: EPA O3 PM. Data derived from Bell et al. (2004).

23.5.2

Morbidity

Associations between ambient air pollutants and respiratory morbidity were examined by Ostro and Rothschild (1989) using the Health Interview Survey (HIS), a large cross-sectional database collected by the National Center for Health Statistics. They attempted to determine the separate health consequences of O3 and particulate matter using six separate years of the HIS. The results, using a fixed effects model that controls for intercity differences, indicate an association between fine PM and both minor restrictions in activity and respiratory

FIGURE 23.10 Comparison of single-day lags (0-, 1-, 2, and 3-day) to a cumulative multiday lag (0–6 day) for percent changes in all cause mortality per 20 ppb increase in 24 year average O3 in all ages. Source: EPA O3CD, (2005). Data derived from Bell et al. (2004).


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TABLE 23.3 Summary of Meta-Analysis of Time-Series Studies Published During the Period 1996–2001

RR < 1

Random Effects Summary Coefficientc

No. of Studies Cause Mortality

Season All

Lag 0

Asthma admissions in children

RR > 1 a

Summer

Any

All

Selected

1h 8h 24 h 1h 24 h 1h 8h 24 h 1h

Selected

8h 24 h 1h

10 (6a) 6 (2a) 3 (0)

1 (0) 1 (0) 1 (0)

0.7 (0.3–1.0) 0.6 (0.2–1.0) 0.1 ( 0.4, 0.6)

8h

4 (2a)

3 (2a)

0.1 ( 1.2, 1.3)

Selectedb

Hospital admissions respiratory

Timing

All

13 (8 ) 9 (5a) 8 (5a) 17 (13a) 22 (8a) 6 (5a) 6 (5a) 2 (2a) 4 (2a)

3 3 3 3 3 1 1 0 1

(0) (0) (1a) (0) (0) (0) (0) (0)

0.2 (0.1–0.3) 0.4 (0.2–0.5) 0.4 (0.1–0.6) 0.3 (0.2–0.4) 0.4 (0.3–0.6) 0.4 (0.1–0.6) 0.6 (0.3–0.9) – 0.5 (0.1–1.0)

a

Number of single studies with a p < 0.05. “selected” lag ¼ If results for more than one lag were presented, the lag selected was chosen as lag focused on by the author, most statistically significant, or largest estimate. c Percentage change per 10 mg/m3 increase and (95% CI), preliminary results. b

conditions severe enough to result in work loss and bed disability in adults. Ozone, however, was associated only with the more minor restrictions. Bates and Sizto (1989) examined associations between ambient air pollutants and hospital admissions for respiratory disease in Southern Ontario. They found a consistent association in summer between hospital admissions for respiratory disease and daily levels of SO42 , O3, and temperature but no association for a group of nonrespiratory conditions. Multiple regression analyses showed that all environmental variables together accounted for 5.6% of the variability in respiratory admissions and that when temperature was forced into the analysis first, it accounted for only 0.89% of the variability. It was found that daily SO42 data collected at one monitoring site in the center of the region were not correlated with respiratory admissions, whereas the SO42 values collected every 6th day, on different days of the week, at 17 stations in the region had the highest correlation with respiratory admissions. They concluded that probably neither O3 nor SO42 alone is responsible for the observed associations with acute respiratory admissions but that either some unmeasured species or some pattern of sequential or cumulative exposure was responsible for the observed morbidity. Burnett et al. (1994) also employed the Ontario acute care hospital database to analyze the effects of air pollution on hospital admissions, but their analysis considered all of Ontario and analyzed the data from each individual hospital, rather than aggregating the counts by region. Slow moving temporal cycles, including seasonal and yearly effects, were removed

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and day-of-week effects were controlled prior to the analysis. Poisson regression techniques were employed because of the low daily admission counts at individual hospitals. O3 displayed a positive association with respiratory admissions in 91% of the 168 hospitals, and 5% of summertime (May through August) respiratory admissions (mean ¼ 107/day) were attributed to O3 (mean ¼ 50 ppb). Positive associations were found in all age groups (0–1, 2– 34, 35–64, and 65þ). A parallel analysis of nonrespiratory admissions showed no such associations. Thurston et al. (1994) focused their analysis of respiratory hospital admissions in the Toronto metropolitan area during the summers (July through August) between 1986 and 1988, when they directly monitored for strong particulate acidity (Hþ) pollution on a daily basis at several sites in that city. Long-wave cycles, and their associated autocorrelations, were removed. Strong and significant positive associations with asthma and respiratory admissions were found for both O3 and Hþ, and somewhat weaker significant associations with SO42 , PM2.5, PM10, and TSP, as measured at a central site in downtown Toronto. No such associations were found either for SO2 or NO2 or for any pollutant with nonrespiratory control admissions. Temperature was only weakly correlated with respiratory admissions and became nonsignificant when entered in regressions with air pollution indices. Simultaneous regressions and sensitivity analyses indicated that O3 was the summertime haze constituent of greatest importance to respiratory and asthma admissions, although elevated Hþ was suggested as a possible potentiator of this effect. During multipollutant, simultaneous regressions on admissions, O3 was consistently the most significant. Of the PM metrics, only Hþ remained statistically significant when entered into the admission regressions simultaneously with O3. Sensitivity analyses also showed that dropping all days with 1 h O3 above 120 ppb (2 of a total 117 days) did not significantly change the O3 coefficients. The simultaneous O3, Hþ, and temperature model indicated that 21 8% of all respiratory admissions during the three summers were associated with O3 air pollution, on average, and that admissions rose an estimated 37 15% above that otherwise expected on the highest O3 day (159 ppb). Moreover, despite differing health care systems, the Toronto regression results for the summer of 1988 were remarkably consistent with previously reported results for that same summer in Buffalo, NY (Thurston et al., 1992). Delfino et al. (1994a) studied daily urgent hospital admissions for respiratory and other illnesses at 31 hospitals in Montreal, Canada, during the warm periods of the year between 1984 and 1988. Both 1 h and 8 h maximum O3 concentrations were considered in the analyses, as well as weather variables (temperature and relative humidity) and PM measurements (Delfino et al., (1994b). For the months of July and August, a significant association was found between all respiratory admissions and both 8 h daily maximum O3 (p 0.01) and 1 h daily maximum O3 (p 0.03) 4 days prior to admission, despite the low O3 concentrations (90th percentile ¼ 60 ppb O3). No significant correlations were found between O3 and nonrespiratory, control admissions. Lipfert and Hammerstrom (1992) reanalyzed the Bates and Sizto (1989) hospital admissions data set for 79 acute-care hospitals in Southern Ontario, incorporating more elaborate statistical methods and extending the data set through 1985. O3, SO42 , and SO2 had significant effects on hospital admissions. By contrast, pollution associations with hospital admissions for accidental causes were nonsignificant in these models. The pollutant mean effect accounted for 19–24% of all summer respiratory admissions. Burnett et al. (1997) extended their study of the effects of O3 on hospitalization for respiratory disease to 16 cities across Canada representing 12.6 million people from 1981 to 1991. There were 720,519 admissions for which the principal diagnosis was a respiratory


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disease. After controlling for SO2, NO2, CO, soiling index, and dew point, the daily high hour concentration of O3 recorded 1 day previous to the date of admission was positively associated with respiratory admissions in the April–December period but not in the winter months. The association between O3 and respiratory hospitalizations varied among cities, with relative risks ranging from 1.000 to 1.088 after simultaneous covariate adjustment. PM and CO were also positively associated with respiratory hospitalizations. Thurston et al. (1992) analyzed admissions to acute-care hospitals in three New York State metropolitan areas during the summers of 1988 and 1989. Environmental variables considered included daily 1 h maximum O3 and 24 h average SO42 and Hþ concentrations, as well as daily maximum temperature recorded at central sites in each community. The strongest O3-respiratory admission associations were found during the period of high pollution in the summer of 1988 and in the most urbanized communities considered (i.e., Buffalo and New York City). After controlling for temperature effects via simultaneous regression, the summer haze pollutants (i.e., SO42 , Hþ, O3) remained significantly related to total respiratory and asthma admissions, but high intercorrelation prevented the clear discrimination of a single pollutant as the causal agent. Depending on the index pollutant, the admission category, and the city considered, it was found that summer haze pollutants accounted for approximately 5–20% of June through August total respiratory and asthma admissions, on average, and that these admissions increased approximately by 30% above average on the highest pollution days. White et al. (1994) reported daily emergency room visit records from June through August 1990 at a large inner city hospital in Atlanta, GA. Daily counts of visits for asthma or reactive airways disease by patients 1–16 years of age (mean ¼ 6.6/day) were related to daily levels of O3, SO2, PM10, pollen, and temperature. The model yielded a 1.42 admissions rate ratio (p ¼ 0.057, 95% CI: 0.99 to 2.0) for the number of asthma visits following days with O3 levels equal to or exceeding a 1 h maximum of 0.11 ppm, which is consistent with the relative risk values reported by Thurston et al. (1992, 1994). In a study of Birmingham, AL, data, Schwartz (1994a) separately examined O3 and PM10 influences on hospital admissions of the elderly for pneumonia (mean ¼ 5.9/day) and COPD (mean ¼ 2.2/day) causes from 1986 to 1989. Base model results (excluding winter months) yielded a 2-day lag RR estimate of 1.14 for pneumonia admissions from a 50 ppb increase in 24 h average O3 (95% CI: 0.94–1.38). Excluding days exceeding 120 ppb yielded similar results (RR ¼ 1.12; CI: 0.92–1.37). For COPD, the basic model yielded RR ¼ 1.17 (CI: 0.86–1.60), whereas excluding days above 120 ppb similarly gave RR ¼ 1.18 (CI: 0.86–1.62). Schwartz (1994b) analyzed O3 and PM10 air pollution relationships with daily hospital admissions of 65-year-old or older persons in the Detroit, MI, metropolitan statistical area from 1986 to 1989. Daily counts for pneumonia (mean ¼ 15.7/day), asthma (mean ¼ 0.75/ day), and all other COPDs (mean ¼ 5.8/day) were regressed on the pollution variables. O3 was analyzed with respect to both its daily 24 h average and 1 h maximum. Both O3 and PM10 were significant in simultaneous pollutant models for pneumonia and COPD but not for asthma (which was ascribed to the low daily counts for this category). According to the regression coefficients and data presented, the mean effect for O3 (11.6%) was double that for PM10 (5.7%) in the pneumonia model, but comparable for COPD (12.2% for O3 versus 10.2% for PM10). Schwartz (1994c) evaluated the associations of both PM10 and O3 with respiratory hospital admissions of the elderly in Minneapolis-St. Paul, MN, from 1986 to 1989. Although no association was found for COPD in the elderly, O3 did make a significant

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independent contribution to hospital admissions of the elderly for pneumonia (mean ¼ 6.0/ day), even after controlling for weather and PM10. In summary, the Schwartz studies of the elderly suggest that a large portion of the O3 effects on total respiratory hospital admissions is contributed by COPD and pneumonia cases in the elderly. According to the results presented by Thurston et al. (1992, 1994, the other major contributor is asthma admissions, which are usually more prevalent in younger age groups. The results of a WHO meta-analysis of hospital admission studies are summarized in Table 23.3. They indicate that ambient O3 often has a significant effect on hospital admissions for respiratory causes (WHO, 2003). A variety of recent population studies have analyzed associations between ambient O3 and emergency room (ER) admissions. Cody et al. (1992) analyzed central New Jersey hospital ER visits to the high O3 season (May through August). For simultaneous regression of respiratory visits on both temperature and O3, there was a significant positive coefficient for O3 and a negative coefficient for temperature. Day-of-week influences were considered but found to be unimportant for these ER visit data. Weisel et al. (1995) examined central New Jersey hospital ER visits for asthma (mean ¼ 5.4/day) during the high O3 season (May through August) for 1986 through 1990. Using a stepwise regression analysis, a significant positive coefficient for O3 and a negative coefficient for morning temperature were found. Other environmental factors considered were not found to be correlated with asthma visits. Stieb et al. (1996) examined the relationship of asthma ER visits to daily concentrations of O3 and other air pollutants in Saint John, New Brunswick, Canada. Data on ER visits with a presenting complaint of asthma (n ¼ 1987) were abstracted for the period 1984– 1992 (May–September). Air pollution variables included O3, SO2, NO2, SO42 , and TSP; weather variables included temperature, dewpoint, and relative humidity. The mean daily 1 h O3 max. concentration during the study period was 41.6 ppb. A positive, statistically significant (p < 0.05), association was observed between O3 and asthma ER visits 2 days later, and the strength of the association was greater in nonlinear models. The frequency of asthma ER visits was 33% higher (95% CI: 10–56%) when the daily 1 h O3 max. concentration exceeded 75 ppb (the 95th percentile). The O3 effect was not significantly influenced by the addition of weather or other pollutant variables into the model or by the exclusion of repeat ER visits. Yang et al. (1997) examined the association between air pollution and the ER visits for asthma in Reno, Nevada, for the period 1992–1994. All three hospitals in the region were included, and there were a total of 1593 ER visits for asthma during this period. The air pollution variables were collected from seven monitoring stations, including PM10, O3, and CO. Levels of pollution were moderately elevated (the average concentrations of PM10, CO, and O3 were 38.0 mg/m3, 4.55 ppm, and 51.0 ppb, respectively). Weighted least-square (WLS) regression and autoregressive integrated moving average (ARIMA) time-series analyses were applied and compared. The daily 1 h maximum O3 concentration was a significant predictor of asthma ER visits. Total asthma visits increased by 33.7% (95% CI: range 6.0–61.5%) for each 100 ppb increase in the O3 level. No association of the concentration of other measured pollutants with daily asthma ER visits was found. Another index of respiratory morbidity that has been studied is clinic visits. HernandezGarduno et al. (1997) monitored patient visits for upper respiratory tract infections in Mexico City at five clinics, and collected data on levels of O3, NO2, CO, SO2, and climatological


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variables. Correlations of filtered data revealed an association between NO2 and O3 and an increase in visits to clinics because of respiratory problems. Autoregressive analysis indicated that pollutant levels/respiratory visits associations remained significant even after simultaneous inclusion of temperature, suggesting that air pollution was associated with 10 to 16% of the clinic visits. High levels of O3 and NO2 increased the total number of clinic visits to 19–43% above average. The other pollutants and the control group did not demonstrate significant associations. Overall, these results are consistent with an O3 effect on asthma morbidity. The results of the recent WHO-EURO meta-analysis on associations between O3 and hospital admissions for respiratory diseases are summarized in Table 23.3. A number of epidemiological studies have shown a consistent relationship between daily variations in ambient oxidant exposure and acute respiratory morbidity in the population. Decreased lung function and increased respiratory symptoms, including exacerbation of asthma, occur with increasing ambient O3, especially in children. Table 23.3 summarizes studies of hospital admissions for asthma in children from the WHO-EURO meta-analysis. In an analysis of respiratory hospital admissions in 14 Canadian cities, Burnett et al. (2001) should that the effect was greatest at a 1- or 2-day lag, but greatest of all for a distributed lag over 4 days. Modifying factors, such as ambient temperature, aeroallergens, and other copollutants (e.g., particles) also can contribute to this relationship. Ozone air pollution can account for a portion of summertime hospital admissions and emergency room visits for respiratory causes. It has been estimated from these studies that O3 may account for roughly 1–3 excess summertime respiratory hospital admissions per 100 ppb O3, per million persons. A recent study by Yang et al. (2003) reported significant associations between O3 respiratory hospital admissions for children less than 3 years of age and for the elderly in Vancouver, Canada, where the 24 h average O3 concentration was only 13.4 ppb. In the Children’s Health Study (CHS), O3 was significantly associated with bronchitic symptoms in children with asthma, but the effects were more strongly associated with organic carbon (OC) and NO2 than with O3 (McConnell et al., 2003). One effect of O3 in the CHS that was not influenced by the other measured pollutants was an increase in absence from school due to respiratory illnesses (Gilliland et al., 2001). Similar effects were seen in a study of asthmatic children in seven United States communities, that is, a strong association of school absence associated most strongly with O3, and respiratory symptoms more strongly associated with pollutants more closely associated with motor vehicle exhaust (O’Connor et al., 2006). Another effect observed in the CHS that was more closely associated with O3 than the other measured pollutants was incident cases of asthma, particularly in children who were engaged in three or more team sports (McConnell et al., 2002). Incident asthma in association with chronic O3 exposure has also been reported in adult males in the AHSMOG cohort study of Seventh Day Adventists in California (McDonnell et al., 1999). The overall implications of the CHS studies on public health and cost–benefit considerations for air pollution control have been reviewed by Kunzli et al. (2003). Another index of morbidity of respiratory morbidity in asthmatics is physicianauthorized medication usage. In their study of children with moderate-to-severe asthma at a summer camp in the Connecticut River Valley, Thurston et al. (2001) found that the camp physician authorized supplemental medication to children in the group at a rate proportional to the ambient O3 concentration.

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23.6 EFFECTS OBSERVED IN STUDIES IN LABORATORY ANIMALS 23.6.1

Effects on Athletic Performance

There are animal models for decreased performance during O3 exposure. Tepper et al. (1985) exposed rats and mice for 6 h to O3 at 80, 120, 250, or 500 ppb while housed in running wheels. Running in both species decreased in a concentration-related manner during exposure to O3, with the decrease being greater with increasing time of exposure. The decrease in running activity produced by O3 persisted for several hours after exposure. At comparable concentrations, activity in rats decreased more than in mice. 23.6.2

Effects on Airways Reactivity

The basis for the effect of O3 on airways reactivity was examined by Gordon et al. (1981) in guinea pigs exposed for 1 h to either 100 or 800 ppb O3. Both exposures significantly inhibited lung cholinesterase activity compared to levels in unexposed animals. The O3-induced responsiveness may be centered on the peripheral lung and be retained long after the O3 exposure ceases, according to a study by Beckett et al. (1988). They exposed the peripheral lungs of anesthetized dogs to 1000 ppb O3 for 2 h using a wedged bronchoscope technique. A contralateral sublobar segment was simultaneously exposed to air as a control. In the O3-exposed segments, collateral resistance (Rcs) was increased within 15 min and remained elevated approximately 150% throughout the 2 h exposure period. Fifteen hours later, the baseline Rcs of the O3-exposed sublobar segments was significantly elevated, and these segments demonstrated increased responsiveness to aerosolized acetylcholine (100 and 500 mg/mL). There were no differences in neutrophils, mononuclear cells, or mast cells (numbers or degree of mast cell degranulation) between O3- and air-exposed airways at 15 h. The small airways of the lung periphery thus are capable of remaining hyperresponsive hours after cessation of localized exposure to O3, but this does not appear to depend on the presence of inflammatory cells in the small airway wall. 23.6.3

Effects on Airway Permeability in Laboratory Tests in Rats

Bhalla et al. (1987) reported that exposure to resting rats to 800 ppb O3 increased tracheal and bronchoalveolar permeability to DTPA at 1 h after the exposure. Bronchoalveolar but not tracheal permeability remained elevated 24 h after the exposure. Exercise during exposure to O3 increased permeability to both tracers in the tracheal and the bronchoalveolar zones and prolonged the duration of increased permeability in the tracheal zone from 1 to 24 h and that in the bronchoalveolar zone from 24 to 48 h. Exposure at rest to 600 ppb O3 plus 2500 ppb NO2 significantly increased bronchoalveolar permeability at 1 and 24 h after exposure, although exposure at rest to 600 ppb O3 alone increased bronchoalveolar permeability only at 1 h after exposure. Exposure to O3 and NO2 during exercise led to significantly greater permeability to DTPA than did exercising exposure to O3 alone. Nitric acid vapor was formed in the O3 þ NO2 atmosphere, suggesting that acidic components in the atmospheres produced effects that were additive on the effect of O3 in producing both increase and prolongation of permeability in tracheal and bronchoalveolar zones of the respiratory tract. Guth et al. (1986) examined changes in apparent lung permeability in rats by measuring the recovery of labeled bovine serum albumin in lung lavage fluid after intravenous injection at the end of the O3 exposure. Their permeability index increased in an exposure

DETERMINANTS OF RESPONSIVENESS TO OZONE EXPOSURES IN ANIMAL STUDIES

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concentration-dependent manner after 6 or 24 h of exposure to O3 at or above levels of 400 ppb. It was also increased after 2 days of exposure to 200 ppb of O3. Abraham et al. (1984) measured changes in airway permeability of tritiated histamine in sheep after exposure to O3. Measurements made 24 h after a 2 h exposure to 500 ppb showed increased permeability. This persistently increased permeability is consistent with the observations of Bhalla et al. (1987) in rats.

23.6.4

Effects on Airway Inflammation

Arsalane et al. (1999) evaluated Clara cell protein (CC16), a 16–17 kDa protein secreted by Clara cells in the bronchial lining fluid of the lung, as a peripheral marker of the integrity of Clara cells and/or of the bronchoalveolar/blood barrier. CC16 was determined in the serum of rats after a single 3 h exposure to 300, 600, or 1,000 ppb of O3. The urinary excretion of the protein was also studied in rats repeatedly exposed to 1,000 ppb O3, 3 h/day, for up to 10 days. The concentrations of CC16 in the lung or trachea homogenates, the lung CC16 mRNA levels, and classical markers of lung injury in BALF were also determined. O3 produced a transient increase of CC16 concentration in serum that reached values that were, on average, 13 times above normal values 2 h after exposure to 1,000 ppb O3. The intravascular leakage of CC16 was dose-dependent and correlated with the extent of lung injury as assessed by the levels of total protein, LDH, and inflammatory cells in BALF. This effect was most likely responsible for the concomitant marked reduction of CC16 concentrations in BALF and lung homogenate, since the CC16 mRNA levels in the lungs were unchanged, and the absolute amounts of CC16 leaking into serum or lost from the respiratory tract were similar. These changes were paralleled by an elevation of the urinary excretion of CC16 resulting from an overloading of the tubular reabsorption process. These results demonstrated the utility of this assay to detect the increased lung epithelial permeability induced by O3. Broeckaert et al. (2000) applied this assay to humans, specifically to cyclists who exercised for 2 h during episodes of photochemical smog, and found that O3 induces an early leakage of lung Clara cell protein. The protein levels increased significantly into the serum from exposure levels as low as 60–84 ppb. These findings confirmed that there is almost no safety margin for the effects of ambient O3 on airway permeability. The assay of CC16 in the serum represents a sensitive noninvasive test allowing the detection of early effects of ambient O3 on the lung epithelial barrier. 23.7 DETERMINANTS OF RESPONSIVENESS TO OZONE EXPOSURES IN ANIMAL STUDIES There is evidently a large genetic component to this responsiveness. Kleeberger (1995) and Kleeberger et al. (1997) have explored the contribution of genetic background to the pathogenesis of airway responses to air pollutants. Using inbred mice strains, Kleeberger (1995) demonstrated a strong genetic influence on responsiveness to O3, and the follow-up study with NO2, another ambient air oxidant, examined the genetic basis for differences in response to the two agents. Kleeberger et al. (1997) determined significant genetic contributions in susceptibility to lung injury and inflammation induced by single and repeated acute exposures to NO2 and whether similar genetic factors control susceptibility to O3. Nine strains of inbred mice (male, 5–6 weeks) were studied. Each was exposed for 3 h to filtered air (controls) or 15 ppm NO2, and cellular inflammation, epithelial injury, and cytotoxicity were

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measured 2, 6, and 24 h thereafter. NO2 exposure caused significant increases in cytotoxicity and lavageable macrophages, epithelial cells, polymorphonuclear leukocytes, and protein in all strains. Interstrain variation in each of these effects indicated that genetic background contributed a significant portion of the variance in responses to this oxidant. Two strains that were differentially susceptible to 3 h exposure to 15 ppm NO2 [C57BL/6J (B6) and C3H/HeJ (C3)] were also exposed for 6 h/day to 10 ppm NO2 on 5 consecutive days. Each of the responses to NO2 was completely adapted after 5 days in resistant C3 mice. Only the lavageable total protein response was adapted in susceptible B6 mice. To determine whether mechanisms of susceptibility to NO2 and O3 were the same, each strain was exposed for 3 h to filtered air or 2 ppm O3 and inflammation was assessed 6 and 24 h thereafter. Strain distribution patterns for responses to each oxidant were not significantly concordant and indicated that susceptibility mechanisms were different. Results of these studies suggest that there is a strong genetic component to NO2 susceptibility that is partially adaptable and significantly different from O3 susceptibility. Studies in laboratory animals have examined the roles of O3 concentration and exposure time in biochemical and cellular responses. Rombout et al. (1989) exposed mice and rats to 380, 750, 1250, and 2000 ppb O3 for 1, 2, 4, and 8 h and measured BAL protein with both daytime and nighttime exposures. Observation times extended from 1 to 54 h. The responses varied with O3 concentration, duration of exposure, time after the start of the exposure, and minute volume, with time of exposure having a greater than proportional influence. For 4and 8 h exposures, the protein content of BAL peaked at 24 h and remained at elevated levels even at 54 h. As indicated previously, Devlin et al. (1991) found increased BAL protein in humans 18 h after an exposure to 100 ppb O3 for 6.6 h. Bhalla and Young (1992) studied the sequence of changes in lung epithelial permeability, free cells in the airways, prostaglandin E2 (PGE2) levels, PMN flux, and alveolar lesions in rats exposed to 800 ppb O3 for 3 h and then studied at 4 h intervals up to 24 h postexposure. Protein content of the BAL increased immediately after O3 exposure and returned to control levels by 16 h postexposure. Albumin concentration in the BAL increased more gradually and the albumin concentrations at 20 and 24 h postexposure were still higher than the control levels. While the total protein in the BAL could be attributed to tissue injury and increased transmucosal transport, the albumin primarily reflected elevated transport from the serum. Total cells in the BAL decreased immediately after the O3 exposure but returned to near normal levels by 4 h. PGE2 levels did not change significantly after O3 exposure. PMNs in the lung sections increased in number with time, peaked at 8 h, and returned to normal levels by 16 h after O3 exposure. The data suggest that the permeability changes may be produced by the direct toxic effects of O3 on the airway epithelia, but the PMNs contribute to the injury process, especially at the later stages. Lung lesions, represented by the thickening of the alveolar septae and increased cellularity, were present at 12 h postexposure and increased with time, thus coinciding with declining permeability at the later stages. The morphological changes lag behind the functional perturbations and appear to represent a phase of functional recovery. The weight of the evidence from these results, showing both functional and biochemical responses in humans and laboratory animals that accumulate over multiple hours and persist for many hours or days after exposure ceases, is clear and compelling. Both functional changes and inflammatory processes were shown to occur in humans following exposures to 100 ppb O3 for 6.6 h, whereas higher concentrations were required to elicit comparable responses in rats. Thus, the rat data, which provide evidence of other effects as well, appear to provide conservative indications of effects on humans.


23.7.1

903

Effect of Single and Multiday Exposures on Particle Clearance

The effects of O3 on mucociliary particle clearance have been studied in rats and rabbits. Rats exposed for 4 h to O3 at concentrations in the range of 400–1200 ppb exhibited a slowing of particle clearance at 800 ppb but not at 400 ppb (Frager et al., 1979; Kenoyer et al., 1981). Rabbits exposed for 2 h to 100, 250, and 600 ppb O3 showed a concentration-dependent trend of reduced clearance rate with increasing concentrations, with the change at 600 ppb being approximately 50% and significantly different from control (Schlesinger and Driscoll, 1987). It is not known why the animal tests show only retarded mucociliary clearance in response to O3 exposure whereas the human tests show accelerated clearance. In corresponding tests with other irritants, that is, H2SO4 aerosol and cigarette smoke, both humans and animals have exhibited accelerated clearance at lower exposures and retarded clearance at higher exposures (Lippmann et al., 1987). Phipps et al. (1986) examined the effects of acute exposure to O3 on some of the factors that affect mucociliary transport rates in studies in which sheep were exposed to 500 ppb O3 for 2 h on two consecutive days. The exposures produced increased basal secretion of sulfated glycoproteins but had no effect on ion fluxes. Histological examination indicated a moderate hypertrophy of submucosal glands in the lower trachea, and they concluded that the exposure caused airway mucus hypersecretion. Studies on the effects of O3 on alveolar macrophage-mediated particle clearance during the first few weeks have also been performed in rats and rabbits. Rats exposed for 4 h to 800 ppb O3 had accelerated particle clearance (Frager et al., 1979). Rabbits exposed to 100, 600, or 1200 ppb O3 once for 2 h had accelerated clearance at 100 ppb and retarded clearance at 1200 ppb. Rabbits exposed for 2 h/day to 100 or 600 ppb O3 for 13 consecutive days had accelerated clearance for the first 10 days, with a greater effect at 600 ppb (Driscoll et al., 1986). The responses of the alveolar macrophages to these exposures were examined by Driscoll et al. (1987). A single exposure to 100 ppb resulted in increased macrophage numbers at 7 days, and repeated exposures resulted in an increase in macrophages and neutrophils on days 7 and 14. Macrophage phagocytosis was depressed immediately and 24 h after acute exposure to 100 ppb and at all times after exposure to 1200 ppb. Repeated exposures to 100 ppb produced reductions in the numbers of phagocytically active macrophages on days 3 and 7, with a return to control levels by day 14. Substrate attachment by macrophages was impaired immediately after exposure to 1200 ppb. The results of these studies demonstrated significant alterations in the numbers and functional properties of alveolar macrophages as a result of single or repeated exposure to 100 ppb O3, a level frequently encountered in areas of high photochemical air pollution. The timescale for the biological integration of the effects of a single O3 exposure has also been examined in studies on laboratory animals. Costa et al. (1989) exposed F-344 rats for 2, 4, and 8 h to O3 at 100, 200, 400 and 800 ppb. Lung function was measured immediately after exposure, and BAL was performed immediately and 24 h later. Functional decrements increased with the product ppb h, leveling off at >6,000, whereas BAL proteins increased rapidly for ppb h >4,000. In another test series involving 6.6 h of exposure with 8% CO2 to stimulate respiration, rats exposed to 500 ppb O3 had functional decrements closely matching those seen in humans at 120 ppb. Thus, rats can provide a good test model for the observed human responses to O3, even though they are a less sensitive species than humans. The lesser responses to a given O3 concentration reported here are consistent with the lesser retention of O3 by rats, as discussed previously.

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This issue was further addressed by Highfill et al. (1992) through an examination of relationships between concentration (C) and exposure time (T) and the impact of changes in the C T product on toxic responses. Using protein concentration of bronchoalveolar lavage fluid as an index of O3-induced lung damage, models were developed from a matrix of C (0.0, 0.1, 0.2, 0.4, and 0.8 ppm) and T (2, 4, and 8 h) values in rat and guinea pig. Equal C T products with different levels of C and T were incorporated into the protocol. Polynomial and exponential least-square models were developed, and the lognormal linear model (Larsen et al., 1991) was evaluated for the rat and guinea pig data. For equal C T products, the results showed similar BALP responses at low C T products. Calculations from the data and the models showed that (1) the models were consistent with reported experiments (Hatch et al., 1986); (2) exercising humans were more responsive to O3 exposure (without adjustments for ventilation rates) than were either rats or guinea pigs as measured by changes in BALP (Koren et al., 1989); and (3) the exponential model provided more generality than Haber’s law by providing estimates of BALP levels for various C T. Ozone-induced bronchial hyperresponsiveness in dogs is inhibited by neutrophil depletion (O’Bryne et al., 1984a) and indomethacin pretreatment (O’Byrne et al., 1984b), suggesting that neutrophils that infiltrate the airways after acute O3 exposure (O’Byrne et al., 1984b; Holtzman et al., 1979) are the cells that release the cyclooxygenase products responsible for O3-induced bronchial hyperreactivity. However, neutrophil infiltration is a relatively late effect (i.e., occurring more than 6 h after exposure) and is not likely to account for the immediate responses. In a follow-up study (Jones et al., 1990), thromboxane antagonists were given to the dogs to further determine the role of thromboxane in O3-induced airway hyperresponsiveness. The antagonists did not inhibit the response, indicating that thromboxane does not have an important role in causing O3-induced airway hyperresponsiveness. Leikauf et al. (1988) investigated the hypothesis that oxidant damage to the tracheal epithelium may result in elaboration of various eicosanoids. After exposure to O3, epithelial cells derived from bovine trachea were isolated and grown to confluency. Monolayers were alternately exposed to O3 and culture medium for 2 h in a specially designed in vitro chamber using a rotating inclined platform (Valentine, 1985). O3 induced increases in cyclooxygenase and lipoxygenase product formation with significant increases in prostaglandins E2, F2a, 6-keto F1a, and leukotriene B4. Release rates of immunoreactive products were dosedependent, and ozone concentrations as low as 100 ppb produced an increase in prostaglandin F2a. Thus, O3 can augment eicosanoid metabolism in airway epithelial cells. 23.7.2

Effects of Single and Multiday Exposures on Lung Infectivity

Both in vivo and in vitro studies have demonstrated that O3 can affect the ability of the immune system to defend against infection. Increased susceptibility to bacterial infection has been reported in mice at 80–100 ppb O3 for a single 3 h exposure (Coffin et al., 1967; Ehrlich et al., 1977; Miller et al., 1978a). Related alterations of the pulmonary defenses caused by short-term exposures to O3 include impaired ability to inactivate bacteria in rabbits and mice (Coffin et al., 1968; Coffin and Gardner, 1972; Goldstein et al., 1977; Ehrlich et al., 1979) and impaired macrophage phagocytic activity, mobility, fragility and membrane alterations, and reduced lysosomal enzymatic activity (Witz et al., 1983; Dowell et al., 1970; Hurst and Coffin, 1971; Hurst et al., 1970; Goldstein et al., 1971a, 1971b; McAllen et al., 1981; Amoruso et al., 1981). Some of these effects have been shown to occur in a variety of species including mice, rats, rabbits, guinea pigs, dogs, sheep, and monkeys.


905

Other studies indicate similar effects for short-term and subchronic exposures of mice to O3 combined with pollutants such as SO2, NO2, H2SO4, and particles (Gardner et al., 1977; Aranyi et al., 1983; Ehrlich, 1980; Grose et al., (1980a, 1980b; Phalen et al., 1980). Similar to the human pulmonary function response to O3, activity levels of mice exposed to O3 have been shown to play a role in determining the lowest effective concentration that alters the immunity (Illing et al., 1980). In addition, the duration of exposure must be considered. In groups of mice exposed to 200 ppb O3 for 1, 3, and 6 h, superoxide anion radical production decreased 8, 18, and 35%, respectively, indicating a progressive decrease in bacteriocidal capacity with increasing duration of exposure (Amoruso and Goldstein, 1988). The major limitation of this large body of data on the influence of inhaled O3 on lung infectivity is that it requires uncertain interspecies extrapolating to estimate the possible effects of O3 on infectivity in humans. Gilmour and Selgrade (1993) compared the pulmonary defenses of rats and mice against streptococcal infection following O3 exposures. In mice, 3 h O3 exposures at 400 ppb resulted in bacterial proliferation and PMN influx in the lungs and excess mortality. By contrast, the rats had only a transient impairment of microbial inactivation. These results indicate that caution is needed in translating the results from either species to predictions of human responses. 23.7.3

Effects of Other Pollutants on Responses to Ozone

A study that addressed the issue of the potentiation of the characteristic functional response to inhaled O3 by other environmental cofactors was performed in Tuxedo, NY (Spektor et al., (1988b). It involved healthy adult nonsmokers engaged in a daily program of outdoor exercise with exposures to an ambient mixture containing low concentrations of acidic aerosols and NO2 as well as O3. Each subject did the same exercise each day, but exercise intensity and duration varied widely between subjects, with minute ventilation ranging from 20 to 153 L, with an average of 79 L, and with duration of daily exercise ranging from 15 to 55 min, with an average of 29 min. Respiratory function was measured immediately before and after each exercise period. The O3 concentrations during exercise ranged from 21 to 124 ppb. All respiratory function indices thus measured showed significant (p < 0.01) O3associated mean decrements. The functional decrements were similar, in proportion to lung volume, to those seen in children engaged in supervised recreational programs in summer camps. They were as large (FEV1) as or much larger (FVC, FEF25–75, PEFR) than those seen in controlled 1- and 2 h exposures in chambers. For the subgroup with the most comparable levels of physical activity, the responses in the field study were even greater. Since the ambient exposures of the adults exercising outdoors were for approximately 1/2 h, compared to the 1- or 2 h exposures in the chamber studies, it was concluded that ambient cofactors potentiate the responses to O3. Thus, the results of the exposures in chambers to O3 in purified air underestimate the O3-associated responses that occur among populations engaged in normal outdoor recreational activity and exposed to O3 in ambient air in the northeastern United States. The apparent potentiation of O3-induced functional decrements observed by Spektor et al. (1988b) in rural New York was not seen by Avol et al. (1984) in a study in southern California in which 42 healthy young men and 8 healthy young women were exposed for 1 h to ambient air containing an average of 153 ppb O3 while exercising heavily in a chamber. The functional decrements were slightly but not significantly smaller than those produced in the same subjects on another day when they were exposed to 160 ppb O3 in purified air. The ambient air in southern California has much higher NO2 concentrations and much lower acid aerosol concentrations than the ambient air in the northeastern United States. Thus, it appears

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that the Hþ content of aerosols is a more likely causal factor for the potentiation seen by Spektor et al. (1988b) than is NO2. However, it must be noted that the Spektor et al. (1988b) study on exercising adults and earlier studies on children at summer camps (Lioy et al., 1985; Spektor et al., (1988a) were not able to demonstrate the specific effect of any of the measured environmental variables, including heat stress and acid aerosol concentration, on the O3-associated responses. The inability to show the individual effects of other environmental cofactors on the response to ambient O3 may result from inadequate knowledge on the appropriate biological averaging time for these other factors. However, in the study of functional responses of children to ambient pollution in Mendham, NJ, a week-long baseline shift in PEFR was associated with both O3 and H2SO4 exposures during a 4-day pollution episode that preceded it (Lioy et al., 1985). A similar response to a brief episode with elevated O3 and a much higher peak 4 h concentration of H2SO4 (46 mg/m3) was seen among girls attending a summer camp in 1986 at Dunnville, on the northeast shore of Lake Erie, Ontario, Canada (Raizenne et al., 1989). Several controlled human exposure studies in chambers have not demonstrated synergy in functional response between O3 and NO2 (Koenig et al., 1988) or between O3 and H2SO4, although Stacy et al. (1983) did report that the mean responses to 400 ppb O3 and 100 mg/m3 H2SO4 after 2 h of exposure at rest were 9.0% for FVC and 11.5% for FEV1, compared to corresponding values of 5.7 and 7.7% for O3 alone, 1.4, and 1.2% for sham exposure, and þ0.9 and þ0.9% for H2SO4 alone. One possible reason why these mean differences, which appear to indicate an enhancement of the O3 response by H2SO4, were not statistically significant was the very high variability of the sham exposure results. By contrast, Koenig et al. (1990) did demonstrate that prior exposure to O3 at 120 ppb with intermittent exercise for 45 min potentiated the subsequent respiratory function response to a 15 min exposure to SO2 at 100 ppb. Frampton et al. (1995) exposed 30 healthy and 30 asthmatic volunteers to either 100 mg/m3 H2SO4 or NaCl for 3 h followed 24 h later by 3 h exposures to O3 at 80, 120, or 180 ppb. For the healthy group, no convincing symptomatic or physiologic effects of exposure to either aerosol or O3 on lung function were found. For the asthmatic group, preexposure to H2SO4 altered the pattern of response to O3 in comparison with NaCl preexposure and appeared to enhance the small mean decrements in FVC that occurred in response to 180 ppb O3. Individual responses among asthmatic subjects were quite variable, some demonstrating reductions in FEV1 of more than 35% following O3 exposure. Analysis of variance of changes in FVC revealed evidence for interactions between aerosol and O3 exposure both immediately after (P ¼ 0.005) and 4 h after (P ¼ 0.030) exposure. Similar effects were seen for FEV1. When normal and asthmatic subjects were combined, four-way analysis of variance revealed an interaction between O3 and aerosol for the entire group (P ¼ 0.0022) and a difference between normal and asthmatic subjects (P ¼ 0.0048). There was no significant effect of exposures on symptoms for either normal or asthmatic subjects. Pollutant interactions that potentiate the characteristic O3 response have also been reported for other effects in controlled exposure studies in animals. Osebold et al. (1980) exposed antigenically sensitized mice to 500 ppb O3 for 3 days, with and without concurrent exposure to 1 mg/m3 of submicrometer H2SO4 droplets. There was an increase in atopic reactivity that was greater than that for each pollutant alone. Lee et al. (1990) exposed 3-month-old male rats to either filtered room air (control) or 1,200 ppb NO2, 300 ppb O3 or a combination of the two oxidants continuously for 3 days. They studied a series of parameters in the lung, including lung weight and enzyme activities related to NADPH generation, sulfhydryl metabolism, and cellular detoxification. The results showed that relative to control, exposure to NO2 caused small (nonsignificant) changes in all the parameters; O3


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caused significant increases in all the parameters except for superoxide dismutase; a combination of NO2 and O3 caused increases in all the parameters, and the increases were greater than those caused by NO2 or O3 alone. Statistical analysis of the data showed that the effects of combined exposure were synergistic for 6-phosphogluconate dehydrogenase, isocitrate dehydrogenase, glutathione reductase, and superoxide dismutase activities, and additive for glutathione peroxidase and disulfide reductase activities but not different from those of O3 exposure for other enzyme activities. Kleinman et al. (1989) reported that lesions in the gas-exhange region of the lung of rats exposed to O3 were greater in size in rats exposed to mixtures containing O3 with either H2SO4 or NO2. Graham et al. (1987) reported a synergistic interaction between O3 and NO2 in terms of mortality in mice challenged with streptococcal infection either immediately or 18 h after pollutant exposure. Last (1989) reported synergistic interaction in rats, in terms of a significant increase in lung protein content, following 9-day exposures at 200 ppb O3 with 20 or 40 mg/m3 H2SO4, and a nonsignificant increase for 9 days at 200 ppb O3 with 5 mg/m3 H2SO4. In summary, single O3 exposures to healthy nonsmoking young adults at concentrations in the range of 80–200 ppb have produced a complex array of pulmonary responses including decreases in respiratory function and athletic performance and increases in symptoms, airway reactivity, neutrophil content in lung lavage, and rate of mucociliary particle clearance. The responses to O3 in purified air in chambers occur at concentrations of 80 or 100 ppb when the exposures involve moderate exercise over 6 h or more and require concentrations of 180 or 200 ppb when the duration of exposure is 2 h or less. On the other hand, mean FEV1 decrements more than 5% have been seen at 100 ppb of O3 in ambient air for children at summer camps and for adults engaged in outdoor exercise for only 1/2 h. The apparently greater responses to O3 in ambient air may result from the presence of, or prior exposures to, acidic aerosol, but this tentative hypothesis need further investigation. Further research is also needed to establish the interrelationships between small transient functional decrements, such as FEV1, PEFR, mucociliary clearance rates, and changes in symptoms, performance, reactivity, permeability, and neutrophil counts. The latter may be adverse or may be more closely related to the accumulation or progression of chronic lung damage. If transient changes in readily measured functions, such as FEV1 or PEFR, are closely correlated with other, more significant health effects, then they could be established as useful surrogates in large-scale laboratory, field, and epidemiologic research as well as further retrospective analyses of published data on human exposure–response. Finally, we need more investigation of the mechanisms underlying the pulmonary responses to inhaled O3. The mechanisms have been summarized by Bates (1989, 1995). The author notes that after O3 exposure, the inspiratory capacity is first reduced as a consequence of a lower maximal negative intrapleural pressure on taking a full inspiration. Maximal inspiratory and expiratory mouth pressures are not affected. He emphasizes that the FEV1/FVC ratio is not initially affected after O3 exposure, which is to say that the FVC and FEV1 initially fall together. He postulates that stimulation of the C-fiber system in the lung must lead to a “braking” effect on the inspiratory muscles as a first consequence of O3 exposure, and this probably occurs as a result of induced inflammation. The increased respiratory rate after O3 exposure, the increased lung permeability, the increased airway reactivity, and the fact that b blockers do not prevent the changes induced by O3 all support this hypothesis. A similar mechanism was suggested by Lee (1990) for the inhibitory effect of the gas phase of cigarette smoke on breathing. He showed that the effect could be largely prevented in rats by the administration of dimethylthiourea, a hydroxyl radical scavenger. The reduced response after repeated

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exposures might result from a thicker lining of mucus over the surface of the airway or from actual cell replacement after exposure.

23.8 EFFECTS OF MULTIPLE DAY AND AMBIENT EPISODE EXPOSURES Since single exposures lasting for 1 h or more at current peak ambient O3 levels produce measurable biological responses in healthy humans and since there is a high probability that one high-O3 day will be followed by several more high-O3 days (Rao, 1988), it is important to know the extent to which the effects accumulate or progress over multiple days. This section reviews the fairly substantial database on functional adaptation to repetitive exposures and the more limited database on biochemical and structural changes that such exposures produce. It should be noted that the data on functional adaptation is largely, but not exclusively, based on studies in human volunteers, whereas the database on biochemical and structural changes caused by O3 is based entirely on studies in laboratory animals. Data on exposures lasting more than 2 weeks are discussed in the Section 23.9. It is well established that repetitive daily exposures, at a level that produces a functional response upon single exposure, result in an enhanced response on day 2, with diminishing responses on days 3 and 4 and virtually no response by day 5 (Farrell et al., 1979; Folinsbee et al., 1980; Hackney et al., 1977). Brookes et al. (1989) found enhanced responses on the second day of successive exposures of exercising young adult males to 350 ppb O3 for 1 h as well as an enhanced response to 250 ppb when the previous day’s exposure was to 350 ppb. In older adults (60–89 years), successive days of 2 h exposures to 450 ppb O3 with light exercise led to small functional decrements on the first 2 days but no changes on successive days (Bedi et al., 1989). For repeated 6.6 h/day exposures to 120 ppb O3, the peak functional response occurs on the first day, with progressively lesser responses after the second, third, and fourth days of exposure. However, for these same subjects, their responsiveness to methocholine challenge peaked on the second day and remained elevated throughout all 5 days of exposure (Folinsbee et al., 1994). The persistent elevation of airway responsiveness is one reason to discount the view of some people that the functional adaptation phenomenon indicates that transient functional decrements are not an important health effect. Additional evidence comes from research in animals showing that persistent damage to lung cells accumulates even as functional adaptation takes place. This kind of functional adaptation to exposure disappears about a week after exposure ceases (Horvath et al., 1981; Kull et al., 1982). The adaptation phenomenon has led some people to conclude that transient functional decrements are not important health effects. However, recent research in animals has shown that persistent damage to lung cells accumulates even as functional adaptation takes place. Tepper et al. (1989) exposed rats to 350, 500, or 1000 ppb O3 for 2.25 h on 5 consecutive days. CO2 (8%) was added to the exposure during alternate 15 min periods to stimulate breathing and thereby increase O3 uptake and distribution. The consequences of exposure on pulmonary function, histology, macrophage phagocytosis, lavageable protein, differential cell counts, and lung tissue antioxidants were assessed. Tidal volume, frequency of breathing, inspiratory time, expiratory time, and maximal tidal flows were affected by O3 during days 1 and 2 at all O3 concentrations. By day 5, these O3 responses were completely adapted at 350 ppb, greatly attenuated at 500 ppb, but showed no signs of adaptation in the group exposed to

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1000 ppb. Unlike the pulmonary function data, light microscopy indicated a pattern of progressive epithelial damage and inflammatory changes associated with the terminal bronchiole region. Over the 5 day testing period, a sustained 37% increase in lavageable protein and 60% suppression of macrophage phagocytic activity were observed with exposure to 500 ppb. There were no changes in differential cell counts. Lung glutathione was initially increased but was within the control range on days 4 and 5. Lung ascorbate was significantly elevated above control on days 3–5. These data suggest that attenuation of the pulmonary functional response occurs while aspects of the tissue response reveal progressive damage. Van Bree et al. (1989) reported the influence of exposure time per day and number of exposure days on biochemical and cellular changes in the lung. Seven-day exposures at 800 ppb produced a loss of normal cilia, a hypertrophic response of Clara cells, and an increase in P-450 isoenzyme activity, whereas 4 day exposures produced increases in protein, G6PDH, and GSHP4. In rats exposed for 2, 4, 8, and 16 days at 400 ppb O3 for 4, 8, and 24 h/ day, the quantity of antioxidant in whole-lung tissue was influenced about twice as much by the exposure duration per day as by the number of exposure days. Finally, in rats exposed to 400 ppb O3 for 12 h at either daytime or nighttime, the effects at night, when they were active, were much greater, once again showing the influence of physical activity on responses to O3. Further indications that functional adaptation, as measured in the days and weeks following exposure, is not fully protective against the development of pathological changes have been provided by a study reported by Farman et al. (1997). This was a follow-up study of one by Last et al. (1993) that showed that rats exposed to 800 ppb O3 and 14.4 ppm NO2 for 6 h daily developed progressive bronchiolitis and pulmonary fibrosis after about 8–10 weeks of exposure, with a high level of mortality. To begin to understand what processes are occurring during the approximately 2- to 2.5-month long period of lesion development, they studied the time course of evolution of fibrotic lesions in rats exposed to O3 and NO2. Rats were sampled weekly for 9 weeks from the onset of exposure, and biochemical and histopathological evaluations were performed. They also quantified the reparative potential of the airway epithelium after 4 and 8 weeks of exposure by in vivo labeling with bromodeoxyuridine (BrdU). Histopathological evaluation indicated a triphasic response temporally: inflammatory and fibrotic changes increased mildly for the first 3 weeks of exposure, stabilized or apparently decreased during 4–6 weeks, and demonstrated severe increases over 7–9 weeks. Biochemical quantification of lung 4-hydroxyproline (collagen) content showed a pattern consistent with the histopathology: no significant differences from controls for the first 3 weeks of exposure, significant increases in collagen content after 4–5 weeks of exposure, and a stabilization of lung collagen content after 6 weeks of exposure. In vivo determination of cumulative labeling indexes showed normal (or slightly decreased) repair of the small airway and alveolar epithelium after 4 weeks of exposure, with significantly diminished reparative capacity after 8 weeks. The diminished reparative capacity of the bronchiolar and alveolar epithelium may be causally linked to the rapid, progressive fibrosis that occurs in this model after about 7–8 weeks of exposure to O3 plus NO2. However, it should be noted that this progression of responses may not be relevant to lower level of exposures, since the Last et al. (1993) paper reported no long-term effects in the rats exposed to 200 ppb O3 plus 3.6 ppm NO2 for 90 days. The effects of multiday O3 exposures of laboratory animals on particle clearance from the lungs and on lung infectivity were reviewed previously. They also show that O3-induced transient effects often become greater with repetitive exposures.

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Effects of a multiday episode-type exposure to O3 on humans in ambient air were described by Lioy et al. (1985). During a study focused on daily variations in lung function among 39 children attending a summer day camp in Mendham, NJ, a summer haze episode occurred in which the daily 1 h peak O3 concentrations exceeded 120 ppb on four consecutive days, with the highest concentration being 185 ppb. During the week following the episode, there were consistent deviations in function from the concentration versus peak flow regressions for the individual children, indicating a persistent loss of lung function during that time. In a subsequent reanalysis of the data from this study, Lioy and Dyba (1989) suggested that the persistence of the reduced function during the week following the episode was more likely due to the cumulative daily exposure than by the daily peak concentrations. In any case, the exposure episode was apparently responsible for an approximately 1-weeklong shift in the function baseline, suggesting that epithelial cell death and regeneration were involved and not just a reflex airway constriction. In summary, successive days of exposure of adult humans in chambers to O3 at current high ambient levels lead to a functional adaptation in that the responses are attenuated by the third day and are negligible by the fifth day. However, a comparable functional adaptation in rats does not prevent the progressive damage to the lung epithelium. Daily exposures of animals also increase other responses in comparison to single exposures, such as a loss of cilia, a hypertrophic response of Clara cells, alterations in macrophage function, and alterations in the rates of particle clearance from the lungs. For children exposed to O3 in ambient air there was a weeklong baseline shift in peak flow following a summer haze exposure of a 4-day duration with daily peak O3 concentrations ranging from 125 to 185 ppb. Since higher concentrations used in adult adaptation studies in chambers did not produce such effects, it is possible that baseline shifts require the presence of other pollutants in the ambient air. A baseline shift in peak flow in camp children was also reported by Raizenne et al. (1989) following a brief episode characterized by a peak O3 concentration of 143 ppb and a peak acidic aerosol concentration of 559 nmol/m ¼ 3.

23.9 CHRONIC EFFECTS OF AMBIENT OZONE EXPOSURES The chronic effects database includes a quite limited amount of information on human effects and a more substantial volume of data on effects seen in laboratory animals undergoing chronic exposures. 23.9.1

Controlled Laboratory Exposure Studies: Human Responses

A study by Linn et al. (1988) in Southern California provided evidence for a seasonal adaptation of lung function. In this study, a group of subjects selected for their relatively high functional responsiveness to O3 had much greater functional decrements following 2 h of exposure to O3 at 180 ppb with intermittent exercise in a chamber in the spring than they did in the following autumn or winter, although their responses in the following spring were equivalent to those in the preceding spring. These findings suggest that some of the variability in acute response coefficients reported for earlier controlled human exposures to O3 in chambers could have been related to seasonal variations in responsiveness, which, in turn, may be related to a long-term adaptation to chronic O3 exposure.

CHRONIC EFFECTS OF AMBIENT OZONE EXPOSURES

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Epidemiological Studies

23.9.2.1 Baseline Respiratory Function Some cross-sectional studies also suggest O3related decrements in respiratory function. Stern et al. (1994) examined differences in the respiratory health status of school children, aged 7–11 years, who resided in 10 rural Canadian communities in areas of moderate and low exposure to regional SO42 and O3 pollution. Five of the communities were located in central Saskatchewan, a low-exposure region and five were located in southwestern Ontario, an area with moderately elevated exposures resulting from long-range atmospheric transport of polluted air masses. Summertime 1 h daily O3 maxima means were 69.0 ppb in Ontario and 36.1 ppb in Saskatchewan. Concentrations of SO42 were three times higher in Ontario than in Saskatchewan; there were no significant differences in levels of PM10 or particulate nitrates. Levels of SO2 and NO2 were low in both regions. After controlling for the effects of age, sex, parental smoking, parental education, and gas cooking, no significant regional differences were observed in symptoms. Children living in Ontario had statistically significant (p < 0.01) mean decrements of 1.7% in FVC and 1.3% in FEV1.0 compared to Saskatchewan children, after adjusting for age, sex, weight, standing height, parental smoking, and gas cooking, but therewere no statistically significant regional differences in the pulmonary flow parameters. The differences could have been due to exposures to either O3 or SO42 or their combination. A more definitive study by Kunzli et al. (1997) regressed mid- and end-expiratory flows (FEF25–75%, FEF75%) against effective exposure to O3. A convenience sample of 130 UC Berkeley freshmen, ages 17–21, participated twice in the same tests (residential history, questionnaire, pulmonary function), 5–7 days apart. Students had to be life-long residents of Northern (SF) or Southern (LA) California. Monthly ambient 8 h O3 concentrations were assigned based on the lifetime residential history and nearby monitoring data for O3. For a 20 ppb increase (interquartile range) in 8 h O3, FEF75% decreased, 14% (95% Cl: 1.0–28.3%) of the population mean FEF75%. The effect on FEF25–75% was 7.2% of the mean. Negative confounding factors were region (SF versus LA), gender, and ethnicity. Lifetime 8 h average O3 ranged from 16 to 74 ppb with little overlap between regions. There was no evidence for different O3 effects across regions. Effects were independent of lifetime mean PM10, NO2, temperature, or humidity. Effects on FEV1 tended to be negative whereas those for FVC, although negative in some models, were inconsistent and small. The strong relationship of lifetime effects of ambient O3 on mid- and end-expiratory flows of college freshmen and the lack of association with FEV1 and FVC are consistent with biologic models of chronic effects of O3 in the small airways. Evidence of chronic effects of O3 were reported by Schwartz (1989) based upon an analysis of pulmonary function data in a national population study in 1976–80, that is, the second National Health and Nutrition Examination Survey (NHANES II). Using ambient O3 data from nearby monitoring sites, he reported a highly significant O3-associated reduction in lung function for people living in areas where the annual average O3 concentrations exceeded 40 ppb. Recent studies of lung function growth in cohorts of children in 12 Southern California communities suggest that ambient air pollutants other than O3 are more responsible for the pulmonary function effects (Gauderman et al., 2000, Gauderman et al., 2002, 2005; Avol et al., 2001). 23.9.2.2 Lung Structure An autopsy study of 107 lungs from 14–25-year-old accident victims in Los Angeles County by Sherwin and Richters (1991) reported that 27% had

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what were judged to be severe degrees of structural abnormalities and bronchiolitis not expected for such young subjects, whereas another 48% of them had similar, but less severe, abnormalities. In the absence of corresponding analyses of lungs of comparable subjects from communities having much lower levels of air pollution, the possible association of the observed abnormalities with chronic O3 exposure remains speculative. Some of the abnormalities observed could have been due to smoking and/ or drug abuse, although the authors noted that published work on the association between smoking and small airway effects showed lesser degrees of abnormality (Cosio et al., 1980). 23.9.2.3 Development of Chronic Disease The effects of chronic exposure to O3 and PM was followed for 10 years in a prospective cohort study by Abbey et al. (1995) in 6340 nonsmoking 7th-Day Adventists living in California. Ambient air monitoring data were available for O3, TSP, SO42 , NO2, and SO2. No significant associations were found for NO2 or SO2. O3 was found significantly associated with increasing severity of asthma and with the development of asthma in males. Measured TSP and SO42 and estimated PM2.5 and PM10 were associated with the development of airway obstructive disease, chronic bronchitis, and asthma, and these were not confounded by the presence of the gaseous pollutants. 23.9.2.4 Effect on Longevity The limited evidence available to date is largely negative. Mendelsohn and Orcutt (1979), in a study utilizing the Public Use Sample containing data on 2 million individuals in the United States obtained both death certificate data and air pollution network data in eight regions of the United States. Highly significant and consistent associations with mortality were found for SO42 . Significant, but weaker and less consistent, associations were seen for SO2 and CO. No significant associations were seen for O3 or NO2. The only other multipollutant studies of annual mortality rates were the six-city study of Dockery et al. (1993) and the American Cancer Society study of Pope et al. (2002). In the six-city prospective cohort study of 8,111 adults over 14–16 years of age, highly significant and consistent mortality effects were seen for PM2.5 and SO42 , with smaller effects indicated for TSP, SO2, and NO2. The variations in O3 across the six cities were too small for effects to be detected. In the Pope et al. (2002) study, there was no significant association of premature mortality with PM2.5, SO42 , and SO2, but not with O3. 23.9.3 Controlled Subchronic and Chronic Laboratory Exposure Studies: Animal Responses Most of the inhaled O3 penetrates beyond the sites in the airways that trigger the functional responses. In this deeper region of the lung, at and just beyond the terminal bronchioles, the effects produced by O3 include changes in biochemical indices, lung inflammation, and airway structure. Furthermore, the effects of O3 exposure in this region appear to be cumulative and persistent, even in animals that have adapted to the exposure in terms of respiratory mechanics (Frank et al., 2001). In a series of inhalation studies, rats were exposed to O3 at constant concentrations of either 120 or 250 ppb for 12 h per day for 6 and 12 weeks, or to a daily cycle with a baseline of 60 ppb for 15 h with a broad peak for 8 h averaging 180 ppb for a period of 3 to 12 weeks.


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Hyperplasia of type I alveolar cells in the proximal alveoli was linearly related to the cumulative O3 dose (Huang et al., 1988). The highest O3 dose is received at the acinus, where the terminal bronchioles lead into alveolar ducts, and a series of studies has shown that the effects of inhaled O3 on lung structure is also greatest in this region. Barry et al. (1985) showed that significant changes occurred in the alveoli just distal to the terminal bronchioles in rats exposed for 12 h/day for 6 or 12 weeks to 120 or 240 ppb O3. In both juvenile and adult rats there were significant increases in the numbers of alveolar type I and type II epithelial cells and alterations in the interstitium and endothelium. From physiological studies of rats that were simultaneously exposed, Raub et al. (1983) reported that there were significant increases in the vital capacity and end-expiratory volume that suggested alterations in distensibility of the lung tissue. For the 6-week exposures at 250 ppb, Barry et al. (1988) reported that exposure to O3 produced alterations in the surface characteristics of ciliated and nonciliated (Clara) cells in both groups. There were significant losses (20–30%) of the surface area contributed by ciliated cells, the luminal surface of Clara cells was decreased by 16–25%, and the number of brush cells per square millimeter of terminal bronchiolar basement membrane was also decreased. Thus, the normal structure of terminal bronchiolar epithelial cells was significantly altered. No statistically significant interaction between the effects of O3 and the animal age at the beginning of the exposure was found. The series of inhalation studies was extended to include tests in which there was a daily cycle with a baseline of 60 ppb for 13 h with a 5-day/week broad peak for 9 h averaging 180 ppb, and containing a 1 h maximum of 250 ppb over a period of 3–12 weeks. Combining the results of these tests with the 6-week studies, Huang et al. (1988) and Chang et al. (1991) reported that hyperplasia of type I alveolar cells in the proximal alveoli was linearly related to the cumulative O3 exposure in the four groups. Thus, there is no threshold for cumulative lung damage. Rats exposed for 6 weeks to clean air or to O3 using the daily cyclic exposure regimen used by Huang et al. (1988) were exposed once for 5 h to an asbestos aerosol by Pinkerton et al. (1989). When they were sacrificed 30 days later, the fiber count in the lungs of the O3exposed animals was three times greater than in the sham-exposed animals. Thus, subchronic O3 exposure can increase the effective dose of insoluble particles, which may have toxic and/or carcinogenic effects. In rats exposed for 12 months by Grose et al. (1989) to the daily exposure cycle used by Huang et al. (1988), an increase in the rate of lung nitrogen washout was observed. Residual volume and total lung capacity were reduced. Glutathione peroxidase and reductase activities were increased, but pulmonary superoxide dismutase remained unchanged. aTocopherol levels were decreased in lung lavage supernatant but were unchanged in lavaged cells, while ascorbic acid and lavage fluid protein were increased. Immunological changes were not observed. Thus, 12 months of exposure to O3 caused (1) functional lung changes indicative of a “stiffer” lung; (2) biochemical changes suggestive of increased antioxidant metabolism; and (3) no observable immunological changes. In a follow-up study in which the same exposure cycle was extended for up to 78 weeks, Tepper et al. (1991) found small, but statistically significant, changes in breathing patterns and mechanisms in unanesthetized, restrained rats at weeks 1, 3, 13, 52, and 78 during postexposure challenge with 0, 4, and 8% CO2. The data indicate that O3 exposure caused an overall increase in expiratory resistance (Rc), but particularly at 78 weeks. The spontaneous frequency of breathing and CO2-induced hyperventilation were also reduced. The decrease

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in frequency depended on a significant increase in the inspiratory time relative to control without a change in the expiratory time. However, light microscopic evaluation of the lung did not reveal any lesions associated with O3 exposure. Chang et al. (1992) extended the analyses of animals exposed for 78 weeks to electron microscopic morphometry. Samples from proximal alveolar regions and terminal bronchioles were obtained by microdissection. Analysis of the proximal alveolar region revealed a biphasic response. Acute tissue reactions after 1 week of exposure included epithelial inflammation, interstitial edema, interstitial cell hypertrophy, and influx of macrophages. These responses subsided after 3 weeks of exposure. Progressive epithelial and interstitial tissue responses developed with prolonged exposure and included epithelial hyperplasia, fibroblast proliferation, and interstitial matrix accumulation. The epithelial responses involved both type I and type II epithelial cells. Alveolar type I cells increased in number, became thicker, and covered a smaller average surface area. These changes persisted throughout the entire exposure period and did not change during the recovery period, indicating the sensitivity of these cells to injury. The main response of type II epithelial cells was cell proliferation. The accumulation of interstitial matrix after chronic exposure consisted of deposition of both increased amounts of basement membrane and collagen fibers. Interstitial matrix accumulation underwent partial recovery during follow-up periods in air; however, the thickening of the basement membrane did not resolve. Analysis of terminal bronchioles showed that short-term exposure to O3 caused a loss of ciliated cells and differentiation of preciliated and Clara cells. The bronchiolar cell population stabilized on continued exposure; however, chronic exposure resulted in structural changes, suggesting injury to both ciliated and Clara cells. Thus, chronic exposure to low levels of O3 caused epithelial inflammation and interstitial fibrosis in the proximal alveolar region and bronchiolar epithelial cell injury. Studies at relatively low O3 concentrations have also been conducted on monkeys. Hyde et al. (1989) exposed them to O3 for 8 h/day for 6 or 90 days to 150 or 300 ppb O3. Responses included ciliated cell necrosis, shortened cilia, and secretory cell hyperplasia with less stored glycoconjugates in the nasal region. Respiratory bronchiolitis observed in 6 days persisted for 90 days of exposure. Even at the lower concentration of 150 ppb O3, nonciliated bronchiolar cells appeared hypertrophied and increased in abundance in respiratory bronchioles. For some chronic effects, intermittent exposures can produce greater effects than those produced by a continuous exposure regime that results in higher cumulative exposures. For example, Tyler et al. (1988) exposed two groups of 7-month-old male monkeys to 250 ppb O3 for 8 h/day either daily or, in the seasonal model, on days of alternate months during a total exposure period of 18 months. A control group breathed only filtered air. Monkeys from the seasonal exposure model, but not those exposed daily, had significantly increased total lung collagen content, chest wall compliance, and inspiratory capacity. All monkeys exposed to O3 had respiratory bronchiolitis with significant increases in related morphometric parameters. The only significant morphometric difference between seasonal and daily groups was in the volume fraction of macrophages. Even though the seasonally exposed monkeys were exposed to the same concentration of O3 for only half as many days, they had larger biochemical and physiological alterations and equivalent morphometric changes as those exposed daily. Lung growth was not completely normal in either exposed group. Thus, longterm effects of oxidant air pollutants that have a seasonal occurrence may be more dependent on the sequence of polluted and clean air than on the total number of days of pollution, and estimations of the risks of human exposure to seasonal air pollutants from effects observed in animals exposed daily may underestimate long-term pulmonary damage. The effects


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observed may be considered directly relevant to human health, especially in view of our knowledge that humans receive even greater local doses of O3 in the vicinity of the acinus than do rats. A number of other interesting chronic exposure studies have been done in animals with O3 concentrations in the range of 300–1000 ppb. Those that appear to provide useful insights into mechanisms of toxic action will also be briefly reviewed. Sherwin and Richters (1985) exposed newborn Swiss–Webster mice to intermittent 300 ppb O3 for 7 h/day, 5 days/week for 6 weeks. O3 exposure increased cell and wall measurements. In contrast to results previously reported for adult animals (Sherwin et al., 1983), there was a greater increase in mean type II cell area than in numbers of type II cells. Effects on the type II cell population implicate damage to the type I alveolar lining cells. The increases in alveolar wall measurements that were found in both the adult and the developing mouse lung imply an alteration of the lung scaffolding and raising the question of impaired regeneration of the epithelial lining. The results of a chronic exposure study in rats by Gross and White (1987) illustrate the importance of exposure pattern on the magnitude of the response. They exposed F-344 rats to 500 ppb O3 for 20 h/day, 7 days/week for 52 weeks and produced only mild functional changes (functional residual capacity, residual volume, and DLco), which returned to normal during 3 months of recovery. Grose et al. (1989), using a 23 h exposure ranging from 60 ppb to a peak 1 h maximum of 250 ppb for 5 days/week, produced comparable functional changes in 1 year. Thus, as in the comparison by Tyler et al. (1988) in monkeys, intermittent exposures, modeled after realistic human exposure conditions, can produce much greater responses per unit dose than continuous exposure at high concentration. These results suggest that the damage results, at least in part, from the repeated attempts to adapt to the irritant challenge as well as to the direct effects of the irritant exposure. To characterize the response of respiratory bronchioles (RBs) to chronic high ambient levels of O3, Moffatt et al. (1987) exposed bonnet monkeys 8 h/day for 90 days to 400 or 640 ppb O3. Significant changes in RB following exposure included (1) a thicker wall and a narrower lumen; (2) a thicker epithelial compartment and a much thicker interstitial compartment; (3) shifts in epithelial cell populations with many more nonciliated bronchiolar epithelial cells and fewer squamous type I epithelial cells; (4) larger nonciliated bronchiolar epithelial cells with a larger compliment of cellular organelles associated with protein synthesis; (5) greater amounts of both interstitial fibers and amorphous ground substance; (6) greater numbers of interstitial smooth muscle cells per epithelial basal lamina surface area; and (7) greater volumes of interstitial smooth muscle, macrophages, mast cells, and neutrophils per epithelial basal lamina surface area. These observations imply that chronic O3 exposure causes a concentration-dependent reactive peribronchiolar inflammatory response and an adaptive response consisting of hypertrophy and hyperplasia of the nonciliated bronchiolar cell. Fujinaka et al. (1985) quantitated the response of RB epithelium and peribronchiolar connective tissue (PCT) to chronic exposure to high ambient levels of O3. Adult male bonnet monkeys were exposed 8 h daily for 1 year to either 640 ppb or filtered air. Significant exposure-related changes were greater volume of RB in the lung, smaller diameter of RB lumen, thicker media and intima of peribronchiolar arterioles, thicker RB epithelium, and thicker PCT. Cellular numerical density increased in cuboidal bronchiolar cells and decreased in type I pneumocytes. Cell volume increases occurred in cuboidal bronchiolar, ciliated, and type 2 cells. PCT changes included more amorphous extracellular matrix, neutrophils, and lymphocytes/plasma cells. It was concluded that centriacinar changes

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caused by long-term exposure to high ambient O3 in bonnet monkeys results in narrowing of RBs primarily by peribronchiolar inflammation (inflammatory cells, fibers, and amorphous matrix) and secondarily through hyperplasia of cuboidal bronchiolar cells. The effects of chronic O3 exposure on lung collagen cross-linking were investigated by Reiser et al. (1987) in two groups of juvenile cynomolgus monkeys exposed to 610 ppb for 8 h/day for 1 year. One group was killed immediately after the exposure period; the second exposed group breathed filtered air for 6 months after the O3 exposure before being killed. Previous studies of these monkeys had revealed that lung collagen content was increased in both exposed groups (Last et al., 1984) in this study. The changes in the group killed at the termination of exposure were characteristic of those seen in lung tissue in the acute stage of experimental pulmonary fibrosis. Although the changes seen in the postexposure group suggest that the lung collagen being synthesized at the time the animals were killed was normal, “abnormal” collagen synthesized during the period of O3 exposure was irreversibly deposited in the lungs. This study suggests that long-term exposure to relatively low levels of O3 may cause irreversible changes in lung collagen structure. Barr et al. (1988) exposed rats to either filtered air or 950 ppb O3 8 h daily for 90 days and examined the centriacinar region of lungs morphologically and morphometrically. After chronic O3 exposure, there was a decrease in terminal bronchiole luminal diameter but no change in total terminal bronchiole volume. The most notable change was a 3.4-fold increase in RB volume. They concluded that RB is formed from the centriacinar alveolar duct. Morphologic parameters supporting this conclusion included the presence of fused basement membrane beneath reactive bronchiolar epithelium in the RB, the presence of similar basal laminar changes in both the RB and the proximal alveolar duct septal tips, and the observation that most severe epithelial damage and inflammation occurred in the most proximal alveolar duct rather than in the terminal bronchiole. The severe injury within the acinus shifts distally as RB segments are formed. Hence, most of the damage occurs at the tips of alveolar septa at the RB alveolar duct junction. The issue of the effects of chronic O3 exposure during childhood on lung development was investigated by Tyler et al. (1987) in studies in 28-day-old rats exposed to filtered air or to 640 or 960 ppb O3, 8 h/night, for 42 nights. A second control group was fed ad libitum and exposed to only filtered air. Half the rats were studied at the end of the 42-night exposures, the rest after a 42-day postexposure period during which all rats were fed ad libitum and breathed filtered air. Rats examined at the end of the exposure period had larger saline and fixed lung volumes. These larger lungs had greater volumes of parenchyma, alveoli, and respiratory bronchioles. Some of these changes persisted throughout a 42-day postexposure period. Thus, O3 inhalation by young rats alters lung growth and development in ways likely to be detrimental, and these changes persist after O3 inhalation stops. In summary, chronic exposures to ambient air appear to produce a functional adaptation that persists for at least a few months after the end of the O3 season but dissipates by the spring. Several population-based studies of lung function indicate that there may be an accelerated aging of the lung associated with living in communities with persistently elevated ambient O3, but the limited ability to accurately assign exposure classifications of the various populations in these studies makes a cautious assessment of these data prudent. The plausibility of accelerated aging of the human lung from chronic O3 exposure is greatly enhanced by the results of a series of recent chronic animal exposure studies in rats and monkeys (especially those in rats of Huang et al. (1988) and Grose et al. (1989) using a daily cycle with a 180 ppb average over 9 h superimposed on a 13 h base of 60 ppb and those in monkeys of Hyde et al. (1989) and Tyler et al. (1988) using 8 h/day of 150 and 250 ppb).

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The persistent cellular and morphometric changes produced by these exposures in the terminal bronchioles and proximal alveolar region and the functional changes consistent with a stiffening of the lung reported by Raub et al. (1983) and Tyler et al. (1988) are certainly consistent with the results of the epidemiological studies.

23.10 AMBIENT AIR QUALITY STANDARDS AND GUIDELINES The U.S. Occupational Safety and Health Administration’s permissible exposure limit (PEL) for O3 is 100 parts per billion (ppb), equivalent to 235 mg/m3, as a time-weighted average for 8 h/day, along with a short-term exposure limit of 300 ppb for 15 min (U.S. DOL, 1989). The American Conference of Governmental Industrial Hygienists (ACGIH, 2008) threshold limit value (TLV) for occupational exposure is 100 ppb as an 8 h time-weighted average for light work; 80 ppb for moderate work; and 50 ppb for heavy work. For a 2 h exposure at any workload, the TVL is 200 ppb. The initial primary (health-based) National Ambient Air Quality Standard, established by the Environmental Protection Agency in 1971, was 80 ppb of total oxidant as a 1 h maximum not to be exceeded more than once per year. The NAAQS was revised in 1979 to 120 ppb of O3 as a 1 h maximum not to be exceeded more than four times in 3 years. This initial revision was based on clinical studies by DeLucia and Adams (1977), showing that exercising asthmatic adults exposed for 1 h to 150 ppb in a test chamber had increased cough, dyspnea, and wheezing, along with small but nonsignificant reductions in pulmonary function (U.S. EPA, 1986). A small margin of safety was applied to protect against adverse effects not yet uncovered by research and effects whose medical significance is a matter of disagreement. EPA initiated a review of the 1979 NAAQS in 1983, completed a criteria document for ozone in 1986, and updated it in 1992 (U.S. EPA, 1992). EPA decided, in March 1993, to maintain the existing standard, but to proceed as rapidly as possible with the next round of review. This expedited review was completed with the publication of both a new criteria document (U.S. EPA, 1996) and staff paper (U.S. EPA, 1996) in 1996. In July 1997, the EPA administrator promulgated a revised primary O3 NAAQS of 80 ppb as an 8 h timeweighted average daily maximum, with no more than four annual exceedances, and averaged over 3 years. The change from one allowable annual exceedance to four was to minimize the designation of NAAQS nonattainment in a community that was triggered by rare meteorological conditions especially conducive to O3 formation. The switch to an 8 h averaging time recognized that ambient O3 in much of the United States has broad daily peaks and that human responses are more closely related to the total daily exposure than to brief peaks of O3 exposure. Since the 120 ppb, 1 h average, one exceedance NAAQS was approximately equivalent to an 8 h average, four exceedance NAAQS at a concentration a bit below 90 ppb in average stringency in the United States as a whole, the 1997 NAAQS represents about a 10% reduction in permissible O3 exposure. On March 12, 2008, the EPA administrator issued revised O3 NAAQS to replace the ones promulgated in 1997. The primary (health based) NAAQS, with an 8-hour averaging time, was reduced to 0.075 ppm, and once again the secondary (welfare-based) NAAQS was made equal to the primary NAAQS. Since the official interpretation of the NAAQS at 0.08 ppm was that an exceedance did not occur unless the concentration was greater than 84 ppb, the specification of the NAAQS to three significant figures meant a tightening of the daily NAAQS limit by about 10%. In promulgating the 2008 O3 NAAQS, the administrator disregarded

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the unanimous recommendations of his Clean Air Scientific Advisory Committee (CASAC) that the primary standard be set at a level between 0.060 to 0.070 ppm and that the secondary standard should be based on a seasonal form that better reflected foliar damage and productivity (Letter Report: EPA-CASAC-08-009). The World Health Organization (WHO) has developed Air Quality Guidelines for O3 to assist member states in establishing their own standards. 23.10.1

Guidelines

The second edition of the WHO AQG set the guideline value for O3 at 120 mg/m3 for an 8 h daily average. Since the mid-1990s, there has been no major addition to the evidence from chamber or field studies. There has, however, been a marked increase in health effects

TABLE 23.4

WHO Ozone Air Quality Guideline and Interim Target Daily Maximum 8 h Mean

High level

3

240 mg/m (120 ppb) 160 mg/m3 (80 ppb)

WHO interim target-1 (IT-1)

WHO Air Quality Guidelines (AQGs)

100 mg/m3 (50 ppb)

Effects at the Selected Ozone Level Significant health effects; substantial proportion of vulnerable population affected. Important health effects, an intermediate target for populations with 03 concentrations above this level. Does not provide adequate protection to public health. Rationale: . Lower level 6.6 h chamber exposures of healthy exercising young adults, who show physiological and inflammatory lung effects. . Ambient level at various summer camp studies showing effects on the health of children. . Estimated 3–5% increase in daily mortalitya (based on findings of daily time-series studies). . This concentration will provide adequate protection to public health, though some health effects may occur below this level. Rationale: Estimated 1–2% increase in daily mortalitya (based on findings of daily time-series studies). . Extrapolation from chamber and field studies based on the likelihood that real-life exposure tends to be repetitive and chamber studies do not study highly sensitive or clinically compromised subjects or children. . Likelihood that ambient O is a marker for 3 related oxidants. .

a Deaths attributable to ozone concentrations above estimated baseline of 70 mg/m3. Based on the range of 0.3–0.5%, increase in daily mortality for an increase of 10 mg/m3 8 h O3.

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evidence from epidemiological time-series studies. Combined evidence from those studies show convincing, though small, positive associations between daily mortality and O3 levels, independent of the effects of particulate matter. Similar associations have been observed in both North America and Europe. These time-series studies have shown effects at O3 concentrations below the previous guideline of 120 mg/m3 (60 ppb) without clear evidence of a threshold. Evidence from both chamber and field studies also indicate that there is a considerable individual variation in response to O3. In view of these considerations, there is a good case for reducing the WHO AQG from the existing level of 120 mg/m3. In 2005, the WHO air quality guidelines for O3 was set at the level of ozone: 100 mg/m3 (50 ppb) for daily maximum 8 h mean (WHO, 2006). See Table 23.4. WHO acknowledged that it is possible that health effects will occur below this level in some sensitive individuals. Based on time-series studies, the number of attributable deaths brought forward can be estimated at 1–2% on days when O3 concentrations reaches this guideline level compared to the background O3 level. There is some evidence that O3 also represents unmeasured toxic oxidants arising from similar sources. Measures to control O3 are also likely to control the effects of these pollutants. Hemispheric background concentrations of tropospheric O3 vary in time and space, but can reach average levels of around 80 mg/m3 (40 ppb). These arise from both anthropogenic and biogenic emissions of O3 precursors and downward intrusion of stratospheric O3 into the troposphere. The proposed guideline value may occasionally be exceeded due to natural causes. There is some evidence that long-term exposure to O3 may have chronic effects, but it is not sufficient to recommend an annual guideline. As concentrations increase above the guideline value, health effects at the population level become increasingly numerous and severe. Such effects can occur in places where concentrations are currently high due to human activities or during episodes of very hot weather. The 8 h interim target-1 level has been set at 160 mg/m3 (80 ppb) at which measurable, though transient, changes in lung function and lung inflammation among some healthy young adults have been shown with intermittent exercise in controlled chamber tests. Although some would argue that these responses may not be adverse effects and that they were seen only with vigorous exercise, these views are counterbalanced by the possibility that there are substantial numbers of persons in the general population, including persons of different ages, pre-existing health status, and coexposures that might be more susceptible than the relatively young and generally healthy subjects who were studied. Furthermore, chamber studies provide little evidence about repeated exposure. The exposure to 160 mg/m3 is also likely to be associated with the same effects noted at 100 mg/m3. Based on time-series evidence, attributable deaths can be estimated at 3–5% for daily exposures above the estimated background. At concentrations exceeding 240 mg/m3 (120 ppb), important health effects are likely. This is based on findings from a large number of clinical inhalation and field studies. Both healthy adults and asthmatics would experience significant reductions in lung function as we all as airway inflammation that would cause symptoms and alter performance. There are additional concerns about increased respiratory morbidity in children. Based on time-series evidence, attributable deaths can be estimated at 5–9% for daily exposures above the estimated background. The WHO guidelines and interim targets for ozone are summarized in Table 27.4.

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23.11 SUMMARY AND CONCLUSIONS The apparently reversible effects that have followed acute exposures lasting from 5 min to 6.6 h include changes in lung capacity, flow resistance, epithelial permeability, and reactivity to bronchoactive challenges. These effects may persist for many hours or days after the exposure ceases. Repetitive daily exposures over several days or weeks can exacerbate and prolong these effects. Most of the data available on transient functional effects of O3 were obtained from controlled human exposure studies and field studies of limited duration. Such studies can provide information on chronic pollutant effects only to the extent that prior exposures affect the transient response to single-exposure challenges. Furthermore, interpretation of the results of such tests is limited by our generally inadequate ability to characterize the nature and/or magnitude of the prior chronic exposures. Most of the limited data we have on the effects of chronic O3 exposures on humans come from epidemiological studies. Epidemiological studies offer the prospect of establishing chronic health effects of longterm O3 exposure in relevant populations and offer the possibility that the analyses can show the influence of other environmental factors on responses to O3 exposure. However, the strengths of any of the associations may be difficult to firmly establish because of the complications introduced by uncontrolled cofactors that may confound or obscure the underlying causal factors. The most convenient and efficient way to study mechanisms and patterns of response to inhaled O3 and of the influence of other pollutants and stresses on these responses is by controlled exposures of laboratory animals. One can study the transient functional responses to acute exposures and establish the interspecies differences in response among different animal species and between them and humans similarly exposed. One can also look for responses that require highly invasive procedures or serial sacrifice and gain information that cannot be obtained from studies on human volunteers. Finally, one can use long-term exposure protocols to study cumulative responses and the pathogenesis of chronic diseases in animals. Other advantages of studies on animals are the ability to examine the presence of and basis for variations in response that are related to age, sex, species, strain, genetic markers, nutrition, the presence of other pollutants, and so on. Among the significant limitations to the use of exposure–response data from animal studies in human risk assessments is our quite limited ability to interpret the responses in relation to likely responses in humans who might be exposed to the same or lower levels. Controlled chronic exposure protocols can be very expensive, limiting the number of variables that can be effectively examined in any given study. For studies focused on the biochemical mechanisms of epithelial cells’ responses to O3, cells can be harvested from humans or animals and exposed to O3 in vitro. Interspecies comparisons of cellular response can often be made, and relatively few animals can provide much study material. However, our ability to interpret the results of in vitro assays in relation to likely effects in humans in vivo is quite limited, even when the studies are done with human cells. The cellular response in vitro may differ from that of the same cells in vivo, and the in vivo controls on cellular metabolism and function, which may play a significant role in the overall response, are absent. In terms of functional effects, we know that single O3 exposures to healthy nonsmoking young adults at concentrations in the range of 80–200 ppb produce a complex array of pulmonary responses including decreases in respiratory function and athletic

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performance and increases in symptoms, airway reactivity, neutrophil content in lung lavage, and rate of mucociliary particle clearance. Responses to O3 in purified air in chambers occur at concentrations of 80 or 100 ppb when the exposures involve moderate exercise over 6 h or more and require concentrations of 180 or 200 ppb when the duration of exposure is 2 h or less. However, mean FEV1 decrements more than 5% have been seen at 100 ppb of O3 in ambient air for children at summer camps and for adults engaged in outdoor exercise for only 1/2 h. The apparently greater responses to O3 in ambient air may be related to the presence of, or prior exposures to, acidic aerosol, but further investigation of this hypothesis is needed. Further research is also needed to establish the interrelationships between small transient functional decrements, such as FEV1, PEFR, and mucociliary clearance rates, which may not, in themselves, be adverse effects, and changes in symptoms, performance, reactivity, permeability and neutrophil counts. The latter may be more closely associated with adversity or in the accumulation or progression of chronic lung damage. Successive days of exposure of adult humans in chambers to O3 at current high ambient levels leads to a functional adaptation in that the responses are attenuated by the third day and are negligible by the fifth day. However, a comparable functional adaptation in rats does not prevent the progressive damage to the lung epithelium. Daily exposures of animals also increase other responses in comparison to single exposures, such as a loss of cilia, a hypertrophic response of Clara cells, alterations in macrophage function, and alterations in the rates of particle clearance from the lungs. The clearest evidence that current ambient levels of O3 are closely associated with health effects in human populations comes from epidemiological studies focused on acute responses. The 1997 revision to the O3 NAAQS relied heavily for its quantitative basis on a study of emergency hospital admissions for asthma in New York City (Thurston et al., 1992) and its consistency with other time-series studies of hospital admissions for respiratory diseases in Toronto, all of Southern Ontario, in Montreal, Canada; and in Detroit and Buffalo in the United States. However, other acute responses, while less firmly established on quantitative bases, are also occurring. Chronic human exposures to ambient air appear to produce a functional adaptation that persists for at least a few months after the end of the O3 season but dissipates by the spring. Several population-based studies of lung function indicate that there may be an accelerated aging of the lung associated with living in communities with persistently elevated ambient O3, but the limited ability to accurately assign exposure classifications of the various populations in these studies makes a cautious assessment of these provocative data prudent. The plausibility of accelerated aging of the human lung from chronic O3 exposure is greatly enhanced by the results of a series of chronic animal exposure studies in rats and monkeys. There is little reason to expect humans to be less sensitive than rats or monkeys. On the contrary, humans have a greater dosage delivered to the respiratory acinus than do rats for the same exposures. Another factor is that the rat and monkey exposures were to confined animals with little opportunity for heavy exercise. Thus, humans who are active outdoors during the warmer months may have greater effective O3 exposures than the test animals. Finally, humans are exposed to O3 in ambient mixtures. The potentiation of the characteristic O3 responses by other ambient air constituents seen in the short-term exposure studies in humans and animals may also contribute to the accumulation of chronic lung damage from long-term exposures to ambient air containing O3.

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The lack of a more definitive database on the chronic effects of ambient O3 exposures on humans is a serious failing that must be addressed with a long-term research program. The potential impacts of such exposures on public health deserve serious scrutiny and, if they turn out to be substantial, strong corrective action. Further controls on ambient O3 exposure will be extraordinarily expensive and will need to be very well justified. In summary, this review has shown that (1) the control of ambient O3 to levels within the current NAAQS presents an intractable problem; (2) the current NAAQS contains little, if any, margin of safety against effects considered to be adverse; and (3) a large fraction of U.S. population resides in communities that exceed the O3 NAAQS. Thus, it is important that health scientists and control agency personnel understand the nature and extent of human exposures and the effects they produce in order to communicate health risks effectively to the public and to help prioritize feasible options for reducing exposures. ACKNOWLEDGMENT This research is part of a Center program supported by Grant ES 00260 from the National Institute of Environmental Health Sciences. REFERENCES Abbey DE, Lebowitz MD, Mills PK, Petersen FF, Beeson WL, Burchette RJ (1995) Long-term ambient concentrations of particulates and oxidants and development of chronic disease in a cohort of nonsmoking California residents. Inhal. Toxicol. 7:19–34. Abraham WM, Delehunt JC, Yerger L, Marchete B, Oliver W Jr (1984) Changes in airway permeability and responsiveness after exposure to ozone. Environ. Res. 34:110–119. ACGIH (2008) Threshold Limit Values and Biological Exposure Indices for 2008. Cincinnati: American Conference of Governmental Industrial Hygienists. Adams WC, Schelegle ES (1983) Ozone and high ventilation effects on pulmonary function and endurance performance. J. Appl. Physiol. Respir. Environ. Exercise Physiol. 55:805–812. Adams WC (2002) Comparison of chamber and face-mask 6.6-hour exposures to ozone on pulmonary function and symptoms responses. Inhal. Toxicol. 14:745–764. Adams WC (2003) Comparison of chamber and face-mask 6.6-hour exposures to 0.08 ppm ozone via square-wave and triangular profiles on pulmonary responses. Inhal. Toxicol. 15:265–281. Adams WC (2006) Comparison of chamber 6.6-hour exposures to 0.04–0.08 ppm ozone via squarewave and triangular profiles on pulmonary responses. Inhal. Toxicol. 18:127–136. Alexis N, Urch B, Tarlo S, Corey P, Pengelly D, O’Byrne P (2000) Cyclooxygenase metabolites play a different role in ozone-induced pulmonary function decline in asthmatics compared to normals. Inhal. Toxicol. 12:1205–1224. Altshuller AP (1977) Eye irritation as an effect of photochemical air pollution. J. Air Pollut. Control Assoc. 27:1125–1126. Altshuller AP (1987) Estimation of the natural background of ozone present at surface rural locations. J. Air Pollut. Control Assoc. 37:1409–1417. Amoruso MA, Goldstein BD (1988) Effect of 1, 3, and 6 hour ozone exposure on alveolar macrophages superoxide production. Toxicologist 8:197 Amoruso MA, Witz G, Goldstein BD (1981) Decreased superoxide anion radical production by rat alveolar macrophages following inhalation of ozone or nitrogen dioxide. Life Sci .12:2215–2221.

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Lippmann M, Lioy PJ, Leikauf G, Green KB, Baxter D, Morandi M, Pasternack B, Fife D, Speizer FE (1983) Effects of ozone on the pulmonary function of children. Adv. Mod. Environ. Toxicol. 5:423–446. Lippmann M, Gearhart JM, Schlesinger RB (1987) Basis for a particle size-selective TLV for sulfuric acid aerosols. Appl. Ind. Hyg. 2:188–199. Makino K, Mizoguchi I (1975) Symptoms caused by photochemical smog. Nippon Koshu Eisei Zasshi 2:421–430. McAllen SJ, Chiu SP, Phalen RF, Rasmussen RE (1981) Effect of in vivo ozone exposure on in vitro pulmonary alveolar macrophage mobility. J. Toxicol. Environ. Health 7:373–381. McDonnell WF, Horstman DH, Hazucha MJ, Seal E, Haak ED, Salaam SA, House DE (1983) Pulmonary effects of ozone exposure during exercise: dose–response characteristics. J. Appl. Physiol. 54:1345–1352. McDonnell WF, Horstman DH, Abdul-Salaam S, House DE (1985a) Reproducibility of individual responses to ozone exposure. Am. Rev. Respir. Dis. 131:36–40. McDonnell WF III, Chapman RS, Leigh MW, Strope GL, Collier AM (1985b) Respiratory responses of vigorously exercising children to 0.12 ppm ozone exposure. Am. Rev. Respir. Dis. 132:875–879. McDonnell WF, Horstman DH, Abdul-Salaam S, Raggio LJ, Green JA (1987) The respiratory responses of subjects with allergic rhinitis to ozone exposure and their relationship to nonspecific airway reactivity. Toxicol. Ind. Health 3:507–517. McDonnell WF, Kehrl HR, Abdul-Salaam S, Ives PJ, Folinsbee LJ, Devlin RB, O’Neil JJ, Horstman DH (1991) Respiratory response of humans exposed to low levels of ozone for 6.6 hours. Arch. Environ. Health 46:145–150. McDonnell WF, Muller KE, Bromberg PA, Shy CM (1993) Predictors of individual differences in acute response to ozone exposure. Am. Rev. Respir. Dis. 147:818–825. McDonnell WF, Stewart PW, Andreoni S, Smith MV (1995) Proportion of moderately exercising individuals responding to low-level, multi-hour ozone exposure. Am. J. Respir. Crit. Care Med. 152:589–596. McDonnell WF, Stewart PW, Andreoni S, Seal E Jr, Kehrl H, Horstman DH, Folinsbee LJ, Smith MV (1997) Prediction of ozone-induced FEV1 changes: effects of concentration, duration, and ventilation. Am. J. Respir. Crit. Care Med. 156:715–722. McDonnell WF, Abbey DE, Nishino N, Lebowitz MD (1999) Long-term ambient ozone concentrations and the incidence of asthma in nonsmoking adults: the AHSMOG study. Environ. Res. 80:110–121. McConnell R, Berhane K, Gilliland F, London SJ, Islam T, Gauderman WJ, Avol E, Margolis HG, Peters JM (2002) Asthma in exercising children exposed to ozone: a cohort study. Lancet 359 (9304):386–391. McConnell R, Berhane K, Gilliland FD, Molitor J, Thomas D, Lurmann F, Avol E, Gauderman WJ, Peters JM (2003) Prospective study of air pollution and bronchitic symptoms in children with asthma. Am. J. Respir. Crit. Care Med. 168 (7):790–797. McKittrick T, Adams WC (1995) Pulmonary function response to equivalent doses of ozone consequent to intermittent and continuous exercise. Arch. Environ. Health 50:153–158. Mendelsohn R, Orcutt G (1979) An empirical analysis of air pollution dose–response curves. J. Environ. Econ. Management 6:85–106. Messineo TD, Adams WC (1990) Ozone inhalation effects in females varying widely in lung size: comparison with males. J. Appl. Physiol. 69:96–103. Miller FJ, Menzel DB, Coffin DL (1978b) Similarity between man and laboratory animals in regional pulmonary deposition of ozone. Environ. Res. 17:84–101. Miller FJ, Illing JW, Gardner DE (1978a) Effects of urban ozone levels on laboratory-induced respiratory infections. Toxicol. Lett. 2:163–169.

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Moffatt RK, Hyde DM, Plopper CG, Tyler WS, Putney LF (1987) Ozone-induced adaptive and reactive cellular changes in respiratory bronchioles of bonnet monkeys. Exp. Lung Res. 12:57–74. National Research Council.(1977) Toxicology. In: Ozone and Other Photochemical Oxidants. Washington, DC: National Academy of Sciences, Committee on Medical and Biologic Effects of Environmental Pollutants. Neas LM, Dockery DW, Koutrakis P, Tollerud DJ, Speizer FE (1995) The association of ambient air pollution with twice daily peak expiratory flow rate measurements in children. Am. J. Epidemiol. 141:111–122. Nightingale JA, Rogers DF, Chung KF, Barnes PJ (2000) No effect of inhaled budesonide on the response to inhaled ozone in normal subjects. Am. J. Respir. Crit. Care Meditor. 161:479–486. O’Bryne PM, Waters EH, Gold BD, Aizawa HA, Fabbri LM, Alpert SE, Nadel JA, Holtzman MJ (1984a) Neutrophil depletion inhibits airway hyperresponsiveness induced by ozone exposure. Am. Rev. Respir. Dis. 130:214–219. O’Byrne PM, Waters EH, Aizawa H, Fabbri LM, Holtzman MJ, Nadel JA (1984b) Indomethacin inhibits the airway hyperresponsiveness but not the neutrophil influx induced by ozone in dogs. Am. Rev. Respir. Dis. 130:220–224. O’Connor GT, Neas L, Vaughn B, Kattan M, Mitchell H, Mastin P, Crain EF, Evans R, Gruchalla R, Morgan W, Stout J, Lippmann M (2006) Acute respiratory health effects of air pollution on asthmatic children in US inner cities. N. Engl. J. Med. (Submitted) Osebold JW, Gerschwin LJ, Zee YC (1980) Studies on the enhancement of allergic lung sensitization by inhalation of ozone and sulfuric acid aerosol. J. Environ. Pathol. Toxicol. 3:221–234. Ostro BD, Rothschild S (1989) Air pollution and acute respiratory morbidity: an observational study of multiple pollutants. Environ. Res. 50:238–247. Ostro BD, Lipsett MJ, Mann JK, Krupnick A, Harrington W (1993) Air pollution and respiratory morbidity in Southern California. Am. J. Epidemiol. 137:691–700. Overton JH, Miller FJ (1987) Modelling ozone absorption in lower respiratory tract. Paper presented at 1987 Annual Meeting of Air Pollution Control Association, New York, June. Overton JH, Graham RC (1989) Predictions of ozone absorption in human lungs from newborn to adult. Health Phys. 57(S1):29–36. Peden DB, Setzer RW Jr, Devlin RB (1995) Ozone exposure has both a priming effect on allergeninduced responses and an intrinsic inflammatory action in the nasal airways of perennially allergic asthmatics. Am. J. Respir. Crit. Care Med. 151:1336–1345. Phalen RF, Kenoyer JL, Crocker TT, McClure TR (1980) Effects of sulfate aerosols in combination with ozone on elimination of tracer particles inhaled by rats. J. Toxicol. Environ. Health 6:797– 810. Phipps RJ, Denas SM, Siekzak MW, Wanner A (1986) Effects of 0.5 ppm ozone on glycoprotein secretion, ion and water fluxes in sheep trachea. J. Appl. Physiol. 60:918–927. Pinkerton KE, Brody AR, Miller FJ, Crapo JD (1989) Exposure to low levels of ozone results in enhanced pulmonary retention of inhaled asbestos fibers. Am. Rev. Respir. Dis. 140:1075–1081. Pope CA III, Burnett RT, Thun MJ, Calle EE, Krewski D, Ito K, Thurston GD (2002) Lung cancer, cardiopulmonary mortality, and long-term exposure to fine particulate matter. JAMA 287:1132– 1141. Raizenne ME, Burnett RT, Stern B, Franklin CA, Spenger JD (1989) Acute lung function responses to ambient acid aerosol exposures in children. Environ. Health Perspect. 79:179–185. Rao ST (1988) Prepared discussion: ozone air quality models. J. Air Pollut. Control Assoc. 38:1129– 1135. Raub JA, Miller FJ, Graham JA (1983) Effects of low level ozone exposure on pulmonary function in adult and neonatal rats. Adv. Mod. Environ. Toxicol. 5:363–367.

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24 PESTICIDES Philip J. Landrigan and Luz Claudio

Synthetic pesticides are a diverse group of chemical compounds, most of them derived from petroleum. Pesticides are used to control insects, unwanted plants, fungi, rodents, and other pests (Hayes and Laws, 1991). Approximately 900 pesticide active ingredients including insecticides, herbicides, rodenticides, and fungicides are currently registered for use (California Department of Pesticide Regulation, 2005). These compounds are mixed with each other and are also blended with “inert” ingredients to produce more than 20,000 commercial pesticide formulations. The United States Environmental Protection Agency (EPA) estimates that in 2001, the most recent year for which data are available, the United States spent $11 billion for over 1.2 billion pounds of pesticide active ingredients (U.S. EPA Office of Prevention, Pesticides and Toxic Substances, 2001). There are also naturally occurring pesticides produced by plants and other organisms; these compounds are discussed in Chapter 20. Pesticides are used in an extraordinarily wide range of settings. In the homes, they control mice, termites, and other rodents. In gardens and lawns as well as along highways and under power-line right-of-ways, pesticides control the growth of unwanted plants. By controlling agricultural pests, pesticides have contributed to dramatic increases in crop yields and in the quantity and variety of the diet (National Research Council, 1993). The agricultural sector is the primary consumer of pesticides, accounting for 76% of use by volume. Industrial, commercial, and governmental users (13%), and home and garden users (11%) account for the remainder (U.S. EPA Office of Prevention, Pesticides and Toxic Substances, 2001). Herbicides for weed control account for the largest volume of agricultural pesticide use (59%) and are applied primarily on corn and soybeans. Insecticides are the next major category of agricultural pesticides (21% of volume) and are used primarily on corn, cotton, and soybeans (Schierow, 1996). Because pesticides create risks as well as benefits, their use poses a perennial problem for public policy and regulation (National Research Council, 1993). Pesticides are specifically designed to be toxic to certain species, and this toxicity is the basis of their utility (Hayes and

Environmental Toxicants, Third Edition Edited by Morton Lippmann Copyright Ó 2009 John Wiley & Sons, Inc.

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Laws, 1991). However, many pesticides are also toxic to species beyond those targeted and have, consequently, caused severe damage to ecosystems. Humans are among the species at risk of such unintended effects, and pesticides have been shown to cause a wide range of adverse effects on human health including acute and chronic injury to the nervous system, lung damage, injury to the reproductive organs, dysfunction of the immune and endocrine systems, birth defects, and cancer. Children are especially susceptible to the effects of pesticides on health (National Research Council, 1993). Recognition of the toxicity of pesticides has stimulated enactment of a vast body of protective laws and regulations in nations around the world. It has also fostered the development of a continuing series of newer and less toxic pesticides, and these actions have helped to control the toxic hazards. The National Academy of Sciences estimates that since 1954, a 90% reduction in the toxicity of pesticides applied to food crops has occurred in the United States (National Research Council, 1996).

24.1 EVOLVING PATTERNS OF PESTICIDE USE The era of modern pesticides began in the nineteenth century, when sulfur compounds were developed as fungicides. In the late nineteenth century, arsenical compounds were introduced to control the insects that attack fruit and vegetable crops, for example, lead arsenate was used widely on apples and grapes. All these substances were acutely toxic (Schuman and Simpson, 1997). In the 1940s, the chlorinated hydrocarbon pesticides, most notably DDT (dichlorodiphenyltrichloroethane), were introduced. For a time, DDT and similar chemicals were used extensively in agriculture and for the control of malaria and other insect-borne diseases. Because these pesticides had little or no immediate toxicity, they were widely hailed and initially believed to be safe (Wargo, 1996). The publication of Rachel Carson’s Silent Spring (1962) brought to the attention of the American public the potential of the chlorinated hydrocarbon pesticides for long-term accumulation and toxicity in the food chain. Carson documented that DDT had caused widespread reproductive failure and near extinction of bald eagles and ospreys, two species that had accumulated large quantities of DDT because of their high position in the food chain. In 1972, the newly created EPA banned DDT in the United States. Today, the principal classes of insecticides used in most industrialized countries are organophosphates, carbamates, and pyrethroids. Unlike the chlorinated hydrocarbons, these compounds are short-lived in the environment and do not bioaccumulate. The organophosphates and carbamates are, however, neurotoxicants, and the human nervous system can be affected by some of these compounds, causing serious acute and chronic toxicity (Blondell, 1997). Other commonly used pesticides in current use are known or suspected to be carcinogens, reproductive toxicants, or toxicants to the endocrine system (Costa, 1997; Zahm et al., 1997). Insecticide use has declined in recent years, reflecting in part the adoption of programs such as integrated pest management (IPM)(Benbrook, 1996). These programs emphasize the use of nonchemical means of pest control to replace and complement pesticide use. Elements of IPM in the home include the cleanup of food residues, the sealing of foundation cracks, and good maintenance.Forexample,theproject“GrowingUpHealthyinEastHarlem”,developedbythe Mount Sinai Children’s Environmental Health and Disease Prevention Research Center in partnership with two neighborhood health centers, illustrated the successful implementation of

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IPM with the reduction of organophosphate use in an urban setting (Brenner et al., 2003). In agriculture, integrated pest management may involve crop rotation and the use of resistant plant strains (Baker et al., 2002). Fungicide use has remained steady and herbicide use has increased substantially, whereas the use of insecticides has begun to decline.

24.2 EXPORT OF HAZARDOUS PESTICIDES The export of highly toxic pesticides manufactured but banned in industrially developed nations to developing countries for use remains a major issue in pesticide regulation. Between, 1997 and 2000, the United States exported nearly 65 million pounds of either forbidden or severely restricted pesticides; 57% of which were shipped to developing nations, while the remaining 43% went to ports in the Netherlands and Belgium (Smith, 2001), presumably for transshipment. While application of highly toxic pesticides continues in many nations and especially in less developed countries, the past three decades has witnessed a concerted effort by the international community to monitor, restrict, and eliminate the use of hazardous compounds. The United Nations Environment Programme has provided an international framework for the management of hazardous chemicals and pesticides through the adoption of the Rotterdam Convention on the Prior Informed Consent Procedure for Certain Hazardous Chemicals and Pesticides in International Trade, and also the Stockholm Convention on Persistent Organic Pollutants (POPs). Adopted in 1998, The Rotterdam Convention covers pesticides and industrial chemicals that have been banned or severely restricted for health or environmental reasons and established the “Prior Informed Consent” (PIC) procedure, which mandates that export of a chemical covered by the Convention can only take place with the prior informed consent of the importing party (United Nations Environment Programme, 1998). The Stockholm Convention, in effect since May 2004, seeks the elimination or restriction of production and use of all intentionally produced POPs and targets 12 persistent organic pollutants including nine pesticides aldrin, chlordane, DDT, dieldrin, endrin, heptachlor, hexachlorobenzene, mirex, and toxaphene (United Nations Environment Programme, 2004). Vigorous debate continues on whether DDT should continue to be used, despite the Stockholm Convention, for malaria control in developing countries (Wikipedia DDT, 2006). The main argument in favor of continued use of DDT is its low economic cost. The main arguments against continuing use are(1) the effects of DDT on the environment; and (2) increasing resistance of mosquitoes to the compound. A further argument against continuing us e of DDT for malaria control is that it may result in contamination of food crops intended for export to developed nations, a situation that may cause these products to be rejected by their intended purchasers (Benbrook, 2002).

24.3 EXPOSURE TO PESTICIDES Pesticide exposure may be percutaneous, by inhalation, or by ingestion. Exposure may occur via the diet, in the workplace, in the yard or home and in the community. In assessing exposure, it is important to understand that persons may simultaneously be exposed to multiple pesticides through several routes and that the effects of these multiple exposures may be additive or even synergistic (National Research Council, 1993) (Table 24.1).

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TABLE 24.1 Pesticide Residues on Fruits and Vegetables Heavily Consumed by Young Children Supermarket Warehouse Data 1990–1992

Food Apples Bananas Broccoli Cantaloupes Carrots Cauliflower Celery Cherries Grapes Green beans Leaf lettuce Oranges Peas Peaches Pears Potatoes Spinach Strawberries Tomatoes Total

Number of Samples

Detected Number With One or More Pesticides

Percent With One or More Pesticides Detected

Number of Different Pesticides Detected

542 368 63 225 252 65 114 90 313 249 201 237 191 246 328 258 163 168 395 4,468

425 134 16 78 125 26 85 72 192 95 136 190 87 194 240 120 88 138 203 2,644

78% 36% 25% 35% 50% 40% 75% 80% 61% 38% 68% 80% 46% 79% 73% 47% 54% 82% 51% 59%

25 9 9 19 12 13 13 13 22 20 22 25 19 20 11 17 19 17 22 81

Source: Environmental Working Group (Wiles and Campbell, 1993). Compiled from U.S. EPA. Office of Planning, Policy, and Evaluation, Pesticide Food Residue Database, Anticipated Pesticide Residues in Food.

24.3.1

Occupational Exposure

Occupational exposure to pesticides occurs among manufacturers and formulators; during transport and storage; among mixers, loaders, and applicators working in fields, greenhouses, parks, and residential buildings; among vector control and structural applicators; and among farm workers entering fields or greenhouse workers handling foliage previously sprayed by pesticides (Blondell, 1997; McConnell, 1994). Crop duster aviation mechanics have also been reported to be at high risk for pesticide poisoning. Other groups occasionally exposed include emergency crews or sewer workers involved in cleanup. In developed countries, a very large exposed group consists of building maintenance workers who apply insecticides in public and private housing, schools, hospitals, and commercial structures. 24.3.2

Environmental Exposure

Environmental exposure to pesticides can occur through consumption of contaminated water, ingestion of pesticide residues in food, inhalation of airborne spray drift, exposure to pesticides applied in the home, school or community, or from exposure to improperly disposed hazardous waste. Heaviest use of pesticides in the home has been found to occur in inner-city neighborhoodsforthecontrolofroachesinapartments.InNewYorkState,heaviestuseofpesticidesinall counties statewide occurred in Manhattan and Brooklyn (Landrigan et al., 1999). Seasonal contamination of drinking water by herbicides is reported each spring in the American Midwest, a pattern that coincides with annual application of these compounds,

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atrazine in particular, prior to spring planting. Although nonoccupational exposure is usually at a low level, numerous episodes of acute illness have resulted from environmental exposure to pesticides (National Research Council, 1993). 24.3.3

Pediatric Exposures

Children are a group at particular risk of exposure to pesticides (National Research Council, 1993). A major route of children’s exposure is through their diet. They may also be exposed to pesticides applied in homes or schools, on lawns and in gardens. Children employed in agriculture or living in migrant farm worker camps are particularly at high risk of exposure to pesticides and of suffering from acute pesticide intoxication (McConnell, 1994). Children’s tissues and organs are rapidly developing, and at various stages in early development these growth processes create windows of great vulnerability to pesticides and other environmental toxicants. An analysis undertaken by the National Academy of Sciences (1993) has established that the unique vulnerability of infants and children to pesticides and other environmental toxicants is based on the following four factors: 24.3.3.1 Children have Greater Exposures to Environmental Toxicants Than Adults Pound for pound of body weight, children drink more water, eat more food, and breathe more air than adults. For example, children in the first 6 months of life consume seven times as much water per unit body weight as does the average American adult. Consequently, children are more heavily exposed to toxicants in air, food, and water than adults. Two behavioral characteristics of infants and children further magnify their exposures: their normal hand-to-mouth activity, and their play close to the ground. And lastly, dermal absorption by young children to certain environmental toxicants such as pesticides is higher than in adults, because of their greater surface area relative to body weight and greater skin permeability. 24.3.3.2 Children’s Metabolic Pathways, Especially in the First Months After Birth, are Immature Compared to Those of Adults In some instances, children are actually better able than adults to cope with environmental toxicants, because they are unable to metabolize toxicants to their active form (Kimmel, 1992). More commonly, however, children are less able to detoxify chemicals such as organophosphate pesticides, and thus are more vulnerable to them (Mortensen et al., 1996; Peto et al., 1991; Bearer, 1995). 24.3.3.3 Infants and Children are Growing and Developing, and Their Delicate Developmental Processes are Easily Disrupted Many organ systems in infants and children, the nervous system in particular, undergo extensive growth and development throughout the prenatal period and the first months and years of extrauterine life. If cells in an infant’s brain are injured by neurotoxic pesticides; or if reproductive development is diverted by endocrine-disrupting pesticides, the resulting dysfunction can be permanent and irreversible (National Research Council, 1993). 24.3.3.4 Because Children have More Future Years of Life Than Most Adults, They have More Time to Develop Chronic Disease That may be Initiated by Early Exposures Exposures sustained early in life, including prenatal exposures, appear more likely to lead to disease than similar exposures encountered later. Also deficits sustained early on may be persistent throughout their lives (National Research Council,

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1993). Early exposures to pesticides have been linked in toxicological studies to the subsequent development of Parkinsonian symptoms (Cory-Slechta et al., 2005). Surveys of foods commonly consumed by children have shown that a high proportion of them contain pesticide residues, and that these foods also frequently contain residues of multiple pesticides. A 2005 study designed to measure dietary organ phosphorus pesticide exposure in children found that the most likely route of exposure to these pesticides was through the diet (Wiles and Campbell, 1993). These observations on children’s exposures to pesticides in food were complemented by a 1995 study that found 16 different pesticides present in some of the baby foods most commonly sold in the U. S. (Wiles and Davies, 1995). These pesticide residues included eight that have been shown to be toxic to the nervous system, five that affect the endocrine system, and eight that are potential carcinogens. Consumption of a diet rich in organically grown fruits and vegetables has been shown to be highly effective in reducing dietary exposure to pesticides. Children who consumed a largely organic diet had dramatically lower levels in urinary pesticide residue levels as compared to classmates who consumed a conventional American diet (Lu et al., 2006). In the past decade, since passage of the Food Quality Protection Act in 1996 (see below), levels of pesticide residues in domestically grown foods in the United States have steadily declined. However, residue levels in imported foods have increased during the same period, with the result that the aggregate U.S. exposure to pesticides in foodstuffs has remained relatively constant. Imported winter fruits, such as table grapes, are the most heavily contaminated imported foodstuffs (Groth et al., 2000).

24.4 EPIDEMIOLOGY OF ACUTE PESTICIDE POISONING Data on pesticide poisonings are sparse, and there is serious underreporting of even acute, life-threatening episodes (Blondell, 1997). The best information on occupational pesticide poisoning in the United States comes from California, where physicians are required by law to report all incidents of pesticide intoxication. In the mid-1990s, the average annual number of occupational pesticide poisoning cases reported in California was approximately 1,500, of which 54% occurred in agriculture (Blondell, 1997). Organophosphates were the class of compounds most frequently involved. By extrapolating California data to the nation, it has been estimated that there are between 10,000 and 20,000 cases of physician-diagnosed pesticide poisoning in the United States per year (U.S. EPA, 1992). Data on nonoccupational pesticide poisonings in the United States are collected by the Consumer Product Safety Commission (CPSC) based on a statistical sample of emergency rooms in 6,000 selected hospitals (Blondell, 1997). In 1990–92, there were an estimated 20,000 emergency room visits resulting from pesticide exposure. Incidence was disproportionately high in children, who accounted for 61% of cases. 24.4.1

Pesticide Epidemiology in the Third World

Pesticide use in developing nations, although rapidly increasing, accounts for only 25% of the 3 million tons of pesticides produced worldwide each year (21). Nonetheless, 90% of the estimated 3 million yearly poisonings worldwide, and more than 99% of the 220,000 deaths, occur in developing countries. Examples of major outbreaks of pesticide-associated disease include an estimated 60,000 illnesses and 2000 to 2500 deaths from exposure to methyl

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isocyanate used as an intermediary in insecticide manufacture in an accidental release in Bhopal, India, and an epidemic of acute parathion poisoning in Jamaica (with 17 deaths) caused by consumption of contaminated, imported wheat flour (Melius, 1998; Diggory et al., 1977). In surveys of agricultural workers in several Asian countries, and of cotton workers in Mexico, 3% to 7% and 13% to 15%, respectively, reported poisoning in the previous year. In addition, pesticide poisoning may be more likely to be underdiagnosed in these nations than in the developed countries (Jeyaratnam, 1990).

24.5 TOXICITY OF PESTICIDES Because the chemistry of pesticides is highly diverse, they are capable of causing a wide range of adverse health effects. Depending on the pesticide, or combination of pesticides, to which an individual or a population is exposed, these effects can involve virtually every organ system in the body. Pesticides can produce acutely toxic effects, delayed effects, and chronic effects. Also, some pesticides are developmental toxicants and others are carcinogens and reproductive toxicants. This review summarizes data on the toxic effects of the major classes of agents. 24.5.1

Insecticides

24.5.1.1 Cholinesterase Inhibitors the carbamates.

This class includes both the organophosphates and

24.5.1.2 Acute Clinical Effects of Organophosphates and Carbamates The toxicity of these two classes of insecticides is similar, and both inhibit neuronal acetylcholinesterase (Costa, 1997). Acute poisonings by organophosphates and carbamates account for the majority of systemic pesticide poisoning cases seen each year in the United States (Blondell, 1997). Inhibition of acetylcholinesterase results in an increase in acetylcholine, with resultant overstimulation of the postsynaptic receptors in the cholinergic nervous system. These effects can be differentiated toxicologically into overstimulation of (1) the central nervous system; (2) the nicotinic receptors (skeletal muscle and autonomic ganglia); and (3) the muscarinic receptors (secretory glands and postganglionic fibers in the parasympathetic nervous system). The nicotinic effects generally appear later, and only in the more severe cases. Headache, anxiety, and sleep disturbance, often accompanied by salivation and anorexia, are common early symptoms of mild overexposure to cholinesterase inhibitors (McConnell, 1994). Chest tightness is a symptom of moderately severe poisoning after dermal exposure, but it can be an early symptom if there is significant exposure via the respiratory tract. Seizures and impaired consciousness occur in the most severe cases. Death can follow respiratory arrest. Other reported acute effects include myopathy (including myocardopathy), hypothermia, liver dysfunction, brady- or tachyarrhythmias, leukocytosis, and acute psychosis. 24.5.1.3 Delayed or Chronic Effects of Organophosphates Chronic low-level exposure to organophosphates may result in weakness and malaise, often accompanied by headache and light-headedness, but without other specific symptoms (McConnell, 1994). Also,

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several additional complications may occur in patients poisoned with organophosphates as follows: 1. Psychosis: There have been case reports of acute and persistent psychotic illness associated with poisoning by organophosphates. 2. Intermediate Syndrome: A recently reported (and rare) so-called intermediate syndrome, characterized by the paralysis of the proximal musculature and muscles of respiration and of cranial motor nerves, occurs one to four days after acute poisoning. Although the mechanism of this intermediate syndrome is not known, it is not thought to result from inhibition of acetylcholinesterase. 3. Organophosphate-induced delayed polyneuropathy (OPIDP): This distal dyingback axonopathy is characterized clinically by cramping muscle pain in the legs, often followed by paresthesia and motor weakness. The onset occurs 10 days to 3 weeks after severe poisoning (Lotti, 1992). There may be marked foot drop and weakness of the distal upper extremities. The pathophysiology of the disease probably requires the inhibition and subsequent aging of a poorly characterized intraneuronal esterase known as neuropathy target esterase (NTE), which is distinct from acetyl cholinesterase (Costa, 1997). 4. Developmental Neurotoxicity: Certain organophosphates, of which chlorpyrifos (CP) has been the most extensively studied, have the ability to cause developmental neurotoxicity when exposures occur during windows of vulnerability during prenatal or early postnatal life, when the nervous system is actively developing and differentiating. Early postnatal exposure of rat pups to low doses of CP has been shown to produce reduction in the number of brain cells, learning deficits, and behavioral difficulties. These functional deficits have been shown to persist into adult life (Slotkin and Seidler, 2007). The developmental toxicity of the organophosphate pesticides appears to be quite disjunct from their systemic toxicity (Slotkin et al., 2006). Parathion, for example, has great systemic toxicity, but relatively little developmental toxicity. Systemic toxicity results principally from cholinesterase inhibition. By contrast, a key pathophysiological mechanism underlying the developmental effects of CP appears to be binding of CP to the nicotinic cholinergic receptor during early brain development (Slikker et al., 2005). Acting through this mechanism, CP has been shown to disrupt the basic cellular machinery that controls the patterns of neural cell maturation and the formation of synapses. These effects do not depend on the inhibition of cholinesterase. These mechanisms of developmental toxicity are likely to be shared by other organophosphates, but the potential developmental toxicity of most of these compounds has not been evaluated in detail (Groth et al., 2000). Recent prospective epidemiological studies of birth cohorts have found human correlates to the developmental neurotoxicity of OPs observed in rodents. Human infants exposed in utero to CP have been found to have smaller head circumference at birth than unexposed babies; this effect is most pronounced in infants born to mothers with low expression levels of the OP-metabolizing enzyme paraoxonase, apparent evidence for a novel gene–environment interaction (Berkowitz et al., 2004). Followup studies of children with biochemically documented exposures to CP in utero have found evidence for slowed reflexes at birth (Young et al., 2005), developmental delays (Engel et al., 2007; Fenster et al., 2007) and increased prevalence of attention deficit/hyperactivity disorder (ADHD) (Rauh et al., 2006).


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24.5.1.4 Pyreththrum and Pyrethroids In the mid-1800s, pyrethrum extracts of chrysanthemum flowers (containing pyrethrin and other active ingredients) were found to be effective insecticides (Hayes and Laws, 1991). They are relatively less toxic than other commonly used insecticides. Pyrethrum may cause asthma or allergic rhinitis and contact dermatitis is frequent. Although pyrethroid insecticides are closely related chemically to the naturally occurring pyrethrin, those in use today are all synthetic. They are more stable in the natural environment than the natural compounds and are used in agriculture and in household pest control. Ingestion of large doses of pyrethroids causes salivation, nausea, vomiting, diarrhea, irritability, tremor, incoordination, seizures, and death. Toxicity is thought to be mediated through delay in the closing of sodium channels after discharge of an action potential, which results in repetitive neuronal discharge (Narahashi, 1985). 24.5.1.5 Organochlorine Insecticides This class of insecticides includes DDT, lindane, and the cyclodienes: aldrin, dieldrin, endrin, and heptachlor. Most uses of these chemically diverse insecticides have been banned or restricted in the developed world because of their environmental persistence and bioaccumulation, two properties that have led to great damage to wildlife (Carson, 1962). Fat levels of DDT in human surveys have decreased markedly since DDTwas banned in the United States in 1973. In many nations of the developing world, concentrations of DDT in human fat and in milk from humans and cows continue to be high. 24.5.1.6 Acute Effects Of Organochlorines Dermal absorption of the cyclodienes, chlordecone, and lindane is high. Most occupational poisonings within this class result from the acute toxicity of chlordecone, endrin, aldrin, and dieldrin (Blondell, 1997). Most other acute poisonings result from the ingestion of these insecticides. Acute toxicity reflects poorly understood neuronal hyperactivity in the central nervous system. Sudden seizures (especially from aldrin, dieldrin, endrin, lindane, and toxaphene) may occur up to 48 h after exposure and may be relatively intractable. Headache, nausea, dizziness, incoordination, confusion, tremor, and paresthesia are common. Tremor is characteristic of poisoning with DDT and chlordecone. Abnormal liver enzyme levels and renal tubular abnormalities may be seen. 24.5.1.7 Chronic Effects of Organochlorines Epidemic occupational chlordecone (Kepone) poisoning in a manufacturing facility in Virginia was characterized by anxiety and tremor, opsoclonus, personality change, oligospermia, pleuritic and joint pains, weight loss, and liver disease. The effects were chronic. In addition, two workers’ wives were poisoned by contact with contaminated work clothes, and a portion of the Chesapeake Bay was polluted by discharge from the plant (Cannon et al., 1978). Almost all organochlorine insecticides have been found to be carcinogenic in at least one species of rodent. Idiosyncratic cases of aplastic anemia have been reported anecdotally in association with exposure to organochlorines, especially to chlordane and lindane (McConnell, 1994). 24.5.2

Herbicides

Herbicides are the most important class of pesticides in terms of U.S. market share. Their use in agriculture and elsewhere is increasing steadily. This diverse class includes atrazine; 2,4-D; 2,4,5-T; glyphosate; and paraquat. Most of these compounds, paraquat excepted, have

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low acute toxicity. Some herbicides, such as alachlor and atrazine, are important groundwater contaminants, and are animal carcinogens. 24.5.2.1 Chlorphenoxy Herbicides Chlorphenoxy herbicides have been used widely in the United States and elsewhere against broad-leaved plants. During the Vietnam War, 2,4-D and 2,4,5-T were applied together, in a 50:50 mixture, by American forces for deforestation and to destroy food crops, in a formulation known as Agent Orange. 2,4,5-T that is no longer marketed in most industrialized countries may become contaminated, during its manufacture, with dioxins, of which the most toxic and intensely studied is the tetrachlorinated 2,3,7,8-tetrachloro-dibenzodioxin (TCDD) isomer. An increased incidence of cancer has been observed in workers chronically and heavily exposed to TCDD. The largest published cohort study of heavily exposed herbicide production workers demonstrated an association between exposure, non-Hodgkin’s lymphoma, and soft tissue sarcoma (Fingerhut et al., 1991). To assess the hazards of exposure to dioxin in Agent Orange among Vietnam War Veterans, the United States National Academy of Sciences convened an expert panel in 1991. After evaluating the available epidemiological and toxicological data, this committee concluded that there was strong evidence for a positive association between dioxin-contaminated herbicide exposure in Vietnam and soft-tissue sarcoma, nonHodgkin’s lymphoma and Hodgkin’s disease. The Committee also concluded that there was strong evidence for a link between Agent Orange and chloracne, as well as for a link with porphyria cutanea tarda. In addition, the Committee found weaker evidence of association between exposure to dioxin-contaminated herbicides and cancer of the lungs, larynx and trachea, as well as prostate cancer and multiple myeloma (National Academy of Sciences, 2006). Further discussion of the toxicity and epidemiology of dioxins and related compounds is presented in Chapter 20. 24.5.2.2 Bipyridils (Diquat and Paraquat) High acute toxicity, lack of an effective antidote, and ready availability (because of low cost and herbicidal efficacy) have contributed to the notoriety of paraquat. 24.5.2.3 Acute Effects of Bipyridils Painfulburnsandbleedingofthegastrointestinaltract are common following acute exposure to paraquat. Approximately 20% of ingested paraquat is absorbed systemically, where it may cause acute hepatic necrosis and renal disease. The most distinctive aspect of paraquat poisoning is delayed pulmonary toxicity. Paraquat is concentrated from the systemic circulation into the lungs, where pulmonary edema may develop two to four days after ingestion. The mortality rate of pulmonary toxicity induced by paraquat is greater than 50% in most case series. The lethal dose in 50% of rabbits dosed dermally is only 4.5 mg/kg/day. Occupational poisoning resulting from dermal absorption is more likely to occur in developing countries, where applicators may carry and mix concentrated paraquat in leaky backpack sprayers. Splashes of paraquat may cause corneal opacification; nosebleeds, rashes, burns, and loss of fingernails are common local effects (McConnell, 1994). 24.5.2.4 Chronic Effects of Bipyridils survivors of acute paraquat poisoning.

Pulmonary fibrosis has been reported among


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Paraquat is structurally similar to 1-methyl-4-phenyl-1,2,3,6-tetrahydropyridine (MPTP), an illicit drug produced as a heroin substitute that caused an outbreak of acuteonset Parkinson’s disease. Although it has been argued that paraquat does not cross the blood–brain barrier, ecologic studies and one case–control study suggest an association between herbicide exposure and Parkinson’s disease (Costa, 1997). Recently, combined exposure in early postnatal life to paraquat and the fungicide Maneb, both of which adversely affect dopamine systems, has been linked in toxicological studies to the development of Parkinsonian symptoms (Cory-Slechta et al., 2005). 24.5.3

Fungicides

Fungicides are applied to seeds, crops, and gardens to prevent growth of fungi. This class of synthetic pesticides encompasses a wide variety of chemicals, including copper, cadmium, organomercury and organotin compounds, substituted benzene, dithiocarbamates, thiophthalimides mancozeb, maneb, pentachlorophenol, and zineb. 24.5.3.1 Dimethyldithiocarbamates The dimethyldithiocarbamates ziram, ferbam, and thiram inhibit acetaldehyde dehydrogenase and have a disulfiram (Antabuse) effect. There have been reports of illness consistent with disulfiram reactions (nausea, vomiting, headache, diaphoresis, thirst, chest pain, and vertigo) among thiram-exposed workers who subsequently consumed ethanol. Ziram and ferbam are irritants and hemolysis occurred in one case of ziram poisoning. All dithiocarbamates are metabolized to carbon disulfide, which may explain similarities in the symptoms of poisoning. 24.5.3.2 Ethylenebisdithiocarbamates The ethylenebisdithiocarbamates (maneb, mancozeb, and zineb) are metabolized to ethylene thiourea, a potent animal carcinogen. This metabolite may also account for the antithyroid effects that occur in animals dosed with these compounds. Ethylene thiourea may concentrate on cooked food previously treated with ethylenebisdithiocarbamates. 24.5.3.3 Alkyl Mercury Alkyl mercury, although little used now in the United States, has resulted in occupational poisoning in seed treating facilities and has produced epidemics of poisoning and death including catastrophic fetotoxicity among farm families in New Mexico and Iraq who consumed treated seed grain or meat from animals that had consumed mercurytreated grain (Pierce et al., 1972; Marsh et al., 1980). Acute poisoning is manifested by headache, metallic taste, paresthesia, tremor, incoordination, slurred speech, constricted visual fields, hearing loss, loss of position sense, and spasticity. Among survivors, permanent neurologic effects are common. 24.5.3.4 Hexachlorobenzene Hexachlorobenzene is a potent inhibitor of uroporphyrinogen decarboxylase, resulting in increases in photosensitive porphyrins. Consumption of hexachlorobenzene-treated seed grain resulted in thousands of cases of poisoning that resembled porphyria cutanea tarda in Turkey in the 1950s. Bullous dermatitis, liver damage, hypertrichosis, and arthritis were permanent in many cases. Hexachlorobenzene levels can be measured in blood and metabolites can be measured in urine. 24.5.3.5 Pentachlorophenol Pentachlorophenol, used to treat lumber, is well absorbed through the skin. It produces an uncoupling of oxidative phosphorylation from oxidative

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metabolism with resultant hyperthermia (McConnell, 1994). Numerous occupational deaths have occurred from overexposure. Symptoms include fever, thirst, sweating, weakness, tachycardia, other arrhythmias and tachypnea. Restlessness, anxiety, and dizziness reflect injury to the central nervous system. 24.5.4

Fumigants and Nematocides

Fumigants and nematocides are a chemically diverse group of pesticides, characterized by vapor pressures that are sufficiently high at room temperature to create airborne concentrations that cause acute toxicity. Inhalation is the principal route of absorption. 24.5.4.1 Acute Effects of Fumigants and central nervous system depression.

Many fumigants cause acute pulmonary edema

24.5.4.2 Methyl Bromide Methyl bromide vapors cause respiratory irritation, pulmonary edema, anorexia, nausea, vomiting, headache, visual disturbances, agitation, dizziness, tremor, incoordination, myoclonus, and muscle weakness. Liquid methyl bromide is absorbed dermally, and causes severe skin burns. Under(1) the Montreal Protocol on Substances that Deplete the Ozone Layer and (2) the Clean Air Act of 1990, production of methyl bromide was to be phased out by January of 2005. However, its production and use still continues under critical use exemptions (CUE). 24.5.4.3 Aluminum Phosphide Aluminum phosphide slowly releases phosphine upon contact with water in air; the release is more rapid in a moist environment. Phosphine is both a mucous membrane- and a respiratory-irritant that produces nausea, vomiting, diarrhea, headache, vertigo, fatigue, paresthesia, cough, dyspnea, chest tightness, and pulmonary edema. Nephro-, hepato-, cardio-, and central nervous system toxicity are common. Patients may smell of rotten fish or garlic. Deaths have occurred among people living near recently fumigated granaries, as a result of early reentry into fumigated structures and aboard grainhauling ships (Wilson et al., 1980). 24.5.4.4 Chronic Effects of Fumigants Survivors of acute methyl bromide poisoning may be left with organic brain damage, seizures, and personality disorders. Ethylene dibromide isa potentanimal carcinogen and,intheoccupational setting,ahumanspermatotoxin. Residues have been found in food and as a fumigant, it is no longer used in the United States. 24.5.4.5 Dibromochloropropane Dibromochloropropane (DBCP) causes decreased sperm counts and testicular atrophy (Babich and Davis, 1981). In 1977, almost one half of a group of poorly protected production workers exposed to DBCP in a plant in California were demonstrated to be azospermic or oligospermia. Recovery was better among initially oligospermic than among azospermic workers (Whorton et al., 1979). DBCP is an animal carcinogen. It has been removed from the continental United States market. 24.5.5

Other Pesticides

24.5.5.1 N,N-diethyltoluamide In the United States, N,N-diethyltoluamide (DEET) is the most commonly used insect repellent. Repellents are purposely applied topically or to

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clothing. Severe irritation of the eyes, irritation of the skin and exacerbation of pre-existing skin disease has been reported. Application of DEET formulations with high concentrations (more than 75% active ingredient) in tropical climates has resulted in severe dermatitis in antecubital and popliteal fossae. It is rapidly absorbed dermally and by ingestion. Behavioral effects have been demonstrated in chronic feeding studies in rats and several idiosyncratic cases of encephalopathy and death have been reported among children heavily treated with DEET. In one study of exposed workers, there was an increased prevalence of symptoms associated with impaired cognitive function and sleep disturbances and a trend toward poorer objectively measured neurobehavioral performance among highly exposed workers (McConnell et al., 1986). 24.5.6

Inert Ingredients

Pesticidal active ingredients represent only a portion of most pesticide formulations. In most countries, the nonpesticide components of commercial pesticide formulations are listed on the label as “inert ingredients.” These materials include solvents, emulsifiers, spreaders, stickers, penetrants, and anticaking agents. The term “inert” is misleading, and reflects only the lack of toxicity of these agents to pests; some inert ingredients are known or suspected to be human toxicants. Volatile mixtures of aliphatic and aromatic hydrocarbons are the most common inert ingredients. Some hydrocarbon agents used to facilitate the penetration of active ingredients to the interior of plants (penetrants) may be skin and eye irritants. The identity of these ingredients, some of which have acute and chronic toxicities, are available in most jurisdictions only at the discretion of the manufacturer (Vacco, 1996). EPA categorizes inert ingredients into four groups: (1) substances known to cause longterm health damage and to harm the environment; (2) those suspected of causing such health and environmental effects; (3) chemicals of unknown toxicity; and (4) those of minimal concern.

24.6 PESTICIDES AND ENDOCRINE/REPRODUCTIVE TOXICITY Concern has arisen in recent years that certain pesticides may have adverse effects on the endocrine system. It has been found that certain organochlorine pesticides such as DDT can interfere with the effects of estrogen. Indeed, it was the study of estrogenic effects of DDT in eagles and ospreys that led to Rachel Carson’s original recognition of the ecotoxicology of the persistent chlorinated hydrocarbon compounds (Carson, 1962). Many pesticides have been found to be estrogenic, including dieldrin, toxaphene, chlordane, DDT, and endosulfan (Soto et al., 1994). The estrogenic and antiestrogenic properties of pesticides may be examined in cell culture models, such as normal human mammary epithelial cells and human breast cancer cells. When these cells are exposed to an estrogenic substance they divide and grow more rapidly, and these effects can be quantified. Because hormones play critical roles in the early development of the immune, nervous, and reproductive systems, even low-dose exposure to endocrine disrupting pesticides during early windows of developmental vulnerability can have devastating effects on fetal life. The developmental effects of exposure to endocrine disrupters will vary depending on age at exposure and sex. It has been proposed that increased exposure to these agents may be the cause of an observed doubling in the incidence of undescended testes in male infants, which

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has been reported since 1960. Effects have also been documented in the sexual development of wildlife in areas where there are elevated levels of pesticides such as DDT in the environment (Guillette et al., 1994). Studies of these effects in humans are in their early stages and are urgently needed. Preliminary analyses suggest associations between reduced sperm counts and elevated levels of pesticides in men’s urine (Swan, 2006). The Food Quality Protection Act of 1996 (see later) now requires that pesticides be tested for potential endocrine toxicity. For further information on endocrine disruptors.

24.7 PESTICIDES AND CHILDHOOD CANCER Steady increases have occurred over the past three decades in the incidence of the two most common pediatric malignancies—leukemia and brain cancer. The cumulative increase in the incidence of primary brain cancer since 1972 has been approximately 40%. During the same time, an increase of more than 50% has occurred in the incidence of testicular cancer among adolescents and young men (National Cancer Institute, SEER Database, 2006; Devesa et al., 1995). To investigate the possible contribution to these trends of prenatal or early postnatal exposures to pesticides, a series of epidemiologic studies have been undertaken. These studies have found consistent, modest associations between early pesticide exposure and childhood cancer. The strongest associations have been found for childhood leukemia and brain cancer. Risk estimates appeared most robust when exposure was characterized in the greatest detail. Highest risks were associated with frequent parental occupational exposure to pesticides and home pesticide use (Daniels et al., 1997). Clearly, more work is required in this area using prospective study designs and precise, real-time measures of exposure.

24.8 LEGISLATIVE FRAMEWORK The regulation of pesticides has been contingent on an assumed need to balance the economic benefits of these compounds against the risks associated with their use (National Research Council, 1993). That is the principle that underlay the Federal Insecticide, Fungicide and Rodenticide Act (FIFRA) of 1947, the first federal statute controlling pesticide use in the United States. This law, which has undergone several revisions, regulated pesticides on the basis that they would not cause “unreasonable adverse effects”. Unreasonable adverse effects were defined in the statute as “any unreasonable risk to man or the environment taking into account the economic, social, and environmental costs and benefits of its use.” This statement implies that some adverse effects were considered reasonable under FIFRA when weighed against potential economic benefits (Schierow, 1996). Due to rising concern about the health effects of pesticides in foods, particularly regarding the potential of some agents to cause cancer, Congress passed the Delaney Clause in 1958, as an amendment to the Federal Food, Drug, and Cosmetic Act. This Clause banned any pesticide that had been shown to cause cancer in humans or animals from processed foods. From the beginning, the Delaney Clause was highly controversial. Representatives of the pesticide and food industries argued that the law was too inflexible as it prohibited processed foods from containing even the smallest amount of carcinogenic chemicals. Environmentalists, in contrast, considered the Delaney Clause a bulwark of public health and environmental protection.

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The Delaney Clause posed a policy dilemma by establishing a standard for pesticides in processed foods much stricter than that established for pesticides in raw foods—this was termed the “Delaney Paradox” (National Research Council, 1987). To copewith this paradox, the EPA for many years did not strictly enforce the Clause and allowed very low, or “de minimis,” levelsof carcinogenic pesticides in processed foods. In the Agency’sopinion, these levels posed no more than a “minimal risk” to health. In 1992, however, the United States Court of Appeals ruled that this approach contravened the intent of the Delaney Clause, and thus was not legal. This decision set the stage for passage in 1996 of the Food Quality Protection Act (FQPA), the major statute regulating pesticides in the United States today. The central premise of the Food Quality Protection Act is that there must be “a reasonable certainty that no harm will result from aggregate exposure to the pesticide chemical residue, including all anticipated dietary exposures and all other exposures for which there is reliable information”. In terms of public health, this is unquestionably a more protective standard than the previous risk–benefit standard based “no unreasonable adverse effect.” The intellectual foundation for the FQPA was provided by a landmark report released in 1993 by the National Academy of Sciences. This report found that the then current laws and regulations did not adequately protect children from the risks of pesticides in foods (National Research Council, 1993). The report recommended that Congress enact a legislation that would specifically consider the effects of pesticide residues on children’s health. Following that guidance, the FQPA requires that . .

. . .

Regulation of pesticides be based on health effects One uniform health-based standard be applied to pesticide residues in both processed and raw foods The latest scientific data be used for the assessment of health risks The regulation of “reduced risk” pesticides be streamlined Pollution prevention be promoted through integrated pest management practices.

Four provisions in the FQPA are aimed specifically at increasing protection for infants and children. Accordingly, the Environmental Protection Agency is directed to 1. Consider children’s sensitivities and unique exposure patterns to pesticides in setting pesticide standards. 2. Explicitly determine that tolerance levels are safe for children. 3. Adopt an additional safety factor of up to tenfold to account for uncertainty in the database relative to children, and to reflect children’s greater exposure and greater susceptibility to pesticides, unless there is reliable evidence that a different factor should be used. Consider sources of pesticide exposure in addition to diet when performing risk assessments and in setting tolerances. Enforcement of the provision of FQPA calling for the imposition of a child protective safety factor, in the setting of pesticide tolerance levels in foods, has been uneven. A “tolerance” is defined as the highest permissible level of a pesticide residue in a particular food. Traditionally, two tenfold safety factors have been employed in setting pesticide tolerance levels. These safety factors are based on the no observed effect level (NOEL), the lowest dose at which effects are seen in toxicological studies of pesticides conducted in

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animals. The first of the two traditional tenfold safety factors is used to account for the extrapolation of toxicological data from animal species to humans. The second accounts for variation among humans. When both are applied in setting a tolerance, the tolerance is set at a level of 1/100th of the NOEL. The new, third, child protective factor was recommended by the National Academy of Sciences committee in their report Pesticides in the Diets of Infants and Children (1993). It is based on the increased biological susceptibility of children as compared to adults and on the substantial differences in exposure that have been demonstrated to exist between children and adults. The third tenfold safety factor is intended in the law as a “default provision”. It is intended to be automatically utilized, unless “reliable data” exist to show that there is no difference in susceptibility between children and adults. To date, the third tenfold safety factor has been applied to the tolerances for only 6 (12%) of 49 organophosphate pesticides that have been examined since the passage of FQPA. A threefold, child protective safety factor has been applied in an additional 6 (12%) instances (Benbrook, 2005b). Thus, for 75% of organophosphate pesticides, despite a widespread lack of data on developmental neurotoxicity, and substantial suspicion that such toxicity exists for many of these compounds, no child protective safety factor has been imposed in setting tolerances (Slotkin, 2004).

24.9 CONCLUSION: ISSUES FOR THE FUTURE Achieving full implementation of the Food Quality Protection Act (FQPA). A major provision of the FQPA is a requirement that a third tenfold, child protective safety factor be applied in setting tolerances for pesticide residues in food. This is meant to protect children against the noncarcinogenic effects of pesticide exposure, especially neurodevelopmental toxicity (Slotkin, 2004). As was noted in the preceding section, enforcement of this requirement has been uneven. Enhanced enforcement will be needed in the future if infants and children are to be adequately protected against the toxic effects of pesticides, especially developmental toxicity. A critical decision now confronting EPA is how to define “reliable data” under FQPA. The traditional toxicologic tests that have been used to assess differences in susceptibility between adult and young animals are relatively insensitive (Tilson, 1995). They tend not to show differences in susceptibility between infants and adults even when such differences may exist, unless the differences lead to gross anatomical defects (Rodier, 1995). Relying on such traditional tests, the EPA failed to apply a third tenfold safety factor in more than twothirds of the first 80 pesticides to come before the Agency for reassessment of tolerance in 1996 and 1997. Concern exists in the environmental community that even chemicals such as lead and PCBs, which are known to be functional neurotoxins but do not cause anatomical birth defects, would be judged “innocent” in this schema. This is an issue that will require close and continuous scrutiny in the years ahead. 24.9.1

Lifetime Toxicity Testing

Too few chemicals are tested for chronic or developmental neurotoxicity (Slotkin, 2004), and those that are examined are typically studied under test protocols in which the chemicals are administered during adolescence and the animals sacrificed and studied 12–24 months later. Functional assessment of neurological function is often not included. This approach

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misses the opportunity to study neurodevelopmental neurotoxicity and to examine possible late effects of early exposures. One approach to overcoming these limitations in design would be to require a lifetime toxicological approach, as has been used by the European Ramazzini Institute for the testing of the carcinogenicity of chemicals (Soffriti et al., 2005). In this approach, duration of toxicity testing protocols would be extended to incorporate administration of chemicals in early life— ideally in utero or even before conception—coupled with lifelong follow-up to natural death. Such expanded protocols may also incorporate functional neurobehavioral test batteries as well as neuropathologic examinations of relevant areas of the brain (Landrigan et al., 2005). 24.9.2

Addressing Pesticide Export Issues

Earlier in this chapter, the concept of the “circle of poison” was discussed. This phenomenon refers to the export of hazardous pesticides from the developed nations, in which they are manufactured but banned, followed by the subsequent return of those highly toxic pesticides on food grown in developing nations and imported to the industrialized world (20). Legislative pressure to resolve this problem has surfaced periodically in the past. To date, this legislative concern has not resulted in the passage of protective legislation. However, it seems likely that renewed legislative concern over this major unresolved issue will rise again. 24.9.3

Ending DDT Manufacture

DDT and other highly persistent and bioaccumulative pesticides continue to be manufactured and used in certain developing nations. A major international effort seeks to bring this manufacture to an end. An economic issue that must be addressed is that the chlorinated hydrocarbon compounds are relatively less expensive to manufacture than the organophosphates, carbamates, and other newer generation pesticides that have been introduced as substitutes. On the other hand, the increasing resistance of insects to DDT and other traditional pesticides may force this transition. 24.9.4

Risk Assessment Versus Pollution Prevention

For the past 20 years, risk assessment has been the predominant paradigm utilized by regulatory agencies to control exposure to pesticides. While risk assessment has had its successes, it is an inherently slow process that typically proceeds by considering only one chemical compound at a time. Moreover, a perennial problem is that pesticides under assessment are presumed innocent until proven to cause injury, and thus they typically remain on the market until the risk assessment is completed. A more effective paradigm for controlling exposures to pesticides consists of pollution prevention. Under this approach, described sometimes as the “precautionary principle”, numerical targets are set for reducing pesticide use over a span of years. One approach to pesticide use reduction is integrated pest management (13).

REFERENCES Babich A, Davis DL (1981) Dibromochloropropane (DBCP): a review. Sci. Total Environ. 17: 207–221.

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Baker BP, Benbrook CM, Benbrook Groth E (2002) Pesticide residues in conventional, integrated pest management (IPM)- grown and organic foods: insights from three US data sets. Food Adit. Contam. 19:427–446. Bearer C (1995) How are children different from adults? Environ. Health Perspect. 103 (Suppl 6): 7–12. Benbrook CM (2005) Tracking the Impacts of the FQPA on Pesticide Dietary Risks: A Preliminary Assessment. 7-7-2005b. Consultant Report to the EPA Office of Inspector General. Berkowitz GS, Wetmur JG, Birman-Deych E, Obel J, Lapinski RH, Godbold JH, Holzman IR, Wolff MS (2004) In utero pesticide exposure, maternal paraoxonase activity, and head circumference. Environ. Health Perspect. 112 (3):388–91. Benbrook C (1996) Pest Management at the Crossroads. Yonkers NY: Consumers Union. Benbrook CM (2002) Organochlorine residues pose surprisingly high dietary risks. J. Epidemiol. Community Health 56:822–823. Blondell J (1997) Epidemiology of pesticide poisonings in the United States, with special reference to occupational cases. Occup. Med. State Art Rev. 12:209–220. Brenner BL, Markowitz S, Rivera M, Romero H, Weeks M, Sanchez E, Deych E, Garg A, Godbold J, Wolff MS, Landrigan PJ, Berkowitz G (2003) Integrated pest management in an urban community: a successful partnership for prevention. Environ. Health Perspect. 111(13):1649–53. California Department of Pesticide Regulation. Newly Registered Active Ingredients. (2005) Available at http://www.cdpr.ca.gov/docs/registration/ais/newreg/2005.pdf Accessed 14 June 2006. Cannon SB, Veazey JM, Jackson RS, Burse VM, Hayes C, Straub WE, Landrigan PJ (1978) Epidemic kepone poisoning in chemical workers. Am. J. Epid. 107:529–537. Carson R (1962) Silent Spring. Cambridge: Riverside Press. Cory-Slechta DA, Thiruchelvam M, Barlow BK, Richfield EK (2005) Developmental pesticide models of the Parkinson disease phenotype. Environ. Health Perspect. 113(9):1263–70. Costa L (1997) Basic Toxicology of Pesticides. Occup. Med. State Art Rev. 12:251–268. Daniels JL, Olshan AF, Savitz D (1997) Pesticdes and childhood cancers. Environ. Health Perspect. 105:1068–1077. DDT chapter. Wikipedia. Accessed 21 July 2006. Devesa SS, Blot WJ, Stone BJ, Miller BA, Tarone RE, Fraumeni JF Jr (1995) Recent cancer trends in the United States. J. Nat. Cancer Inst. 87:175–182. Diggory P, Landrigan PJ, Latimer KP, Ellington AC, Kimbrough RD, Liddle JA, Cline AF, Smrek AL (1977) Fatal parathion poisoning caused by contamination of flour in international commerce. Am. J. Epidemiol. 106:145–153. Engel SM, Berkowitz GS, Barr DB, Teitelbaum SL, Siskind J, Meisel SJ, Wetmur JG, Wolff MS (2007) Prenatal organophosphate metabolite and organochlorine levels and performance on the Brazelton neonatal behavioral assessment scale in a multiethnic pregnancy cohort. Am. J. Epidemiol. Fenster L, Eskenazi B, Anderson M, Bradman A, Hubbard A, Barr DB (2007) In utero exposure to DDT and performance on the Brazelton neonatal behavioral assessment scale. Neurotoxicology. Fingerhut MA, Halperin WE, Marlow DA (1991) Cancer mortality in workers exposed to 2,3,7,8tetrachlorodibenzo-p-dioxin. N. Engl. J. Med. 324:212–218. Groth E, Benbrook CM, Benbrook KL (2000) Pesticide Residues in Children’s Food. Yonkers, NY: Consumers Union. Guillette LJ Jr, Gross TS, Masson GR, Matter JM, Percival HF, Woodward AR (1994) Developmental abnormalities of the gonad and abnormal sex hormone concentrations in juvenile alligators from contaminated and control lakes in Florida. Environ. Health Perspect. 102: 680–688. Hayes WJ, Laws ER (1991) Handbook of Pesticide Toxicology. San Diego: Academic Press.

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Jeyaratnam J (1990) Acute pesticide poisoning: a major global health problem. World Health Stat. Q 43:139–144. Kimmel C (1992) Animal models for assessing developmental toxicity. Similarities and Differences Between Children and Adults. Washington, DC: ILSI Press. pp.43–65. Landrigan PJ, Claudio L, Markowitz SB, Berkowitz GS, Brenner BL, Romero H, Wetmur JG, Matte TD, Gore AC, Godbold JH, Wolff MS (1999) Pesticides and inner-city children: exposures, risks, and prevention. Environ. Health Perspect. (Suppl 3):431–437. Landrigan PJ, Sonawane B, Butler RN, Trasande L, Callan R, Droller D (2005) Early environmental origins of neurodegenerative disease in later life. Environ. Health Perspect. 113:1230–1233. Lotti M (1992) The pathogenesis of organophosphate neuropathy. Crit. Rev. Toxicol. 21:465–487. Lu C, Toepel K, Irish R, Fenske RA, Barr DB, Bravo R (2006) Organic diets significantly lower children’s dietary exposure to organophosphorus pesticides. Environ. Health Perspect. 114 (2):260–263. Marsh DO, Myers GJ, Clarkson TW, Amin-Zaki L, Tikriti S, Majud M (1980) Fetal methyl mercury poisoning: clinical and toxicological data on 29 cases. Ann. Neurol. 7:348–353 McConnell R (1994) Pesticides and related compounds. In: Rosenstock L, Cullen M, editors. Textbook of Clinical Occupational and Environmental Medicine. Philadelphia: WB Saunders Co. McConnell R, Fidler A, Chrislip D (1986) Health Hazard Evaluation Report. Everglades National Park, Everglades, Florida. HETA 83-085-1757. Cincinnati, Ohio: National Institute for Occupational Safety and Health. Melius JM (1998)The Bhopal disaster. In: Rom WN, editor. Environmental and Occupational Medicine. 3rd edn Boston: Little, Brown. Mortensen SR, Chanda SM, Hooper MJ, Padilla (1996) Maturational differences in chlorpyrifosoxonase activity may contribute to age-related sensitivity to chlorpyrifos. J. Biochem. Toxicol. 11:279–287. Narahashi T (1985) Nerve membrane ionic channels as the primary target of pyrethroids. Neurotoxicology 2:3–22. National Academy of Sciences (2006) Health Risks from Dioxin and Related Compounds: Evaluation of the EPA Reassessment. Washington: The National Academies Press. National Cancer Institute. SEER Database.http://seer.cancer.gov/. Accessed 14 February 2006. National Research Council (1996) Carcinogens and Anticarcinogens in the Diet. Washington: National Academy Press. National Research Council (1993) Pesticides in the Diets of Infants and Children. Washington: National Academy Press. National Research Council (1987) The Delaney Paradox. Washington: National Academy Press. Peto R, Gray R, Brantom P, Grasso P (1991) Dose and time relationships for tumor induction in the liver and esophagus of 4080 inbred rats by chronic ingestion of N-nitrosodiethylamine or N-nitrosodimethinalmine. Cancer Res. 51:6452–6469. Pierce PE, Thompson JF, Likosky WH (1972) Alkyl mercury poisoning in humans: a report of an outbreak. JAMA 220:1439–1442. Rauh VA, Garfinkel R, Perera FP, Andrews HF, Hoepner L, Barr DB, Whitehead R, Tang D, Whyatt RW (2006) Impact of prenatal chlorpyrifos exposure on neurodevelopment in the first 3 years of life among inner-city children. Pediatrics 118: (6). Rodier PM (1995) Developing brain as a target of toxicity. Environ. Health Perspect. 103:73–76. Schierow LJ (1996) Pesticide Policy Issues. Washington: The Library of Congress, Congressional Research Service, Environment and Natural Resources Policy Division (IB95016). Schuman SH, Simpson WM Jr (1997) A clinical historical overview of pesticide health issues. Occup. Med. State Art Rev. 12:203–207.

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Slikker W, Xu ZA, Levin ED, Slotkin TA (2005) Mode of action: disruption of brain cell replication, second messenger, and neurotransmitter systems during development leading to cognitive dysfunction—developmental neurotoxicity of nicotine. Crit. Rev. Toxicology 35:703–711. Slotkin TA, Levin ED, Seidler FJ (2006) Comparative developmental neurotoxicity of organophosphate insecticides: effects on brain development are separable from systemic toxicity. Environ. Health Perspect. 114 (5):746–51. Slotkin TA (2004) Guidelines for developmental neurotoxicity and their impact on organophosphate pesticides: a personal view from an academic perspective. Neurotoxicology 25:631–640. Slotkin TA, Seidler FJ (2007) Prenatal chlorpyrifos exposure elicits presynaptic serotonergic and dopaminergic hyperactivity at adolescence: critical periods for regional and sex-selective effects. Reprod Toxicol. 23(3):421–7. Smith C (2001) Pesticide exports from U.S. ports. 1997–2000. Int. J. Occup. Environ. Health. 7 (4):266–274. Soffriti M, Belpoggi F, Degli Espositi D, Lambertini L (2005) Aspartame induces lymphomas and leukaemias in rats. Eur. J. Oncol. 10:107–116. Soto AM, Chung KL, Sonnenschein C (1994) The pesticides endosulfan, toxaphene, and dieldrin have estrogenic effects on human estrogen-sensitive cells. Environ. Health Perspect. 102:380–383. Swan SH (2006) Does our environment affect our fertility? Some examples to help reframe the question. Semin. Reprod. Med. 24:142–146. Tilson HA (1995) The concern for developmental neurotoxicology: is it justified and what is being done about it?. Environ Health Perspect 103(Suppl 6):147–151. U.S. EPA Office of Prevention, Pesticides, and Toxic Substances. Pesticides Industry Sales and Usage: 2000 and 2001 Market Estimates. Available at http://www.epa.gov/oppbead1/pestsales/01pestsales/market_estimates2001.pdf. Accessed 18 June 2006. U.S. Environmental Protection Agency. (1992) Regulatory Impact Analysis of Worker Protection Standards for Agricultural Pesticides. Washington: EPA. United Nations Environment Programme. Division of Technology, Industry and Economics. PIC Rotterdam Convention. Available at http://www.pic.int/index.html Accessed: 10 July 2006. United Nations Environment Programme. Stockholm Convention on Persistent Organic Pollutants (POP). Available at http://www.pops.int/documents/background/hcwc.pdf Accessed: 10 July 2006. Vacco DC (1996) The Secret Hazards of Pesticides: Inert Ingredients. Albany, NY: New York Department of Law, Environmental Protection Bureau. Wargo J (1996) Our Children’s Toxic Legacy. New Haven:Yale University Press. Whorton D, Milby TH, Krauss RM, Stubbs HA (1979) Testicular function in DBCP exposed pesticide workers. J. Occup. Med. 21:161–166. Wiles R, Davies E (1995) Pesticides in Baby Food. Washington:Environmental Working Group. Wiles R, Campbell C (1993) Pesticides in Children’s Food. Washington:Environmental Working Group. Wilson R, Lovejoy FH, Jaeger RJ, Landrigan PJ (1980) Acute phosphine poisoning aboard a grain freighter: epidemiologic, clinical, and pathological findings. JAMA 244:148–1150. Young JG, Eskenazi B, Gladstone EA, Bradman A, Pedersen L, Johnson C, Barr DB, Furlong CE, Holland NT (2005) Association between in utero organophosphate pesticide exposure and abnormal reflexes in neonates. Neurotoxicology. 26 (2):199–209. Zahm SH, Ward MH, Blair A (1997) Pesticides and cancer. Occup. Med. State Art Rev. 12:269–289.

25 SULFUR OXIDES—SO2, H2SO4, NH4HSO4, AND (NH4)2SO4 Morton Lippmann

Avariety of gaseous and particulate chemicals in the ambient air are acidic, and some of them are known to have produced health effects in exposed populations. The best documented of the effects in laboratory studies are attributable to strong acids in aerosol form, that is, sulfuric acid (H2SO4) and ammonium bisulfate (NH4HSO4). Among the weak acids, sulfur dioxide (SO2) and its hydrolysis products have been associated with both acute bronchoconstriction and elevated morbidity and mortality rates. With respect to these elevated rates, SO2 may have been a surrogate for its oxidation products and other particulate matter (PM) pollutants, which often coexist with SO2. This chapter summarizes and discusses the health effects attributable to SO2 and acidic aerosols at concentrations that have been monitored or measured in community and industrial atmospheres. The role of acidic aerosols in the health effects associated with ambient air PM is also discussed in Chapter 10. This chapter does not discuss effects associated with occupational exposures to these sulfur oxide or other acidic pollutants at much higher concentrations because of their lack of relevance to subject at issue, that is, the health effects of atmospheric acidity.

25.1 SOURCES AND EXPOSURES 25.1.1

Sources of Sulfur Oxides

Most of the sulfur in fossil fuel is converted into SO2 in the combustion zone, and it is vented to the atmosphere with the other products of combustion. A small fraction of the sulfur, generally less than 10%, is emitted as H2SO4, with some of it forming a surface film on ultrafine-sized mineral ash particles. When the discharge point is a tall stack, most of the SO2 escapes local deposition on terrestrial surfaces and is gradually (1–10%/h) converted into SO3, a highly hygroscopic vapor. The SO3 rapidly combines with water vapor to Environmental Toxicants, Third Edition Edited by Morton Lippmann Copyright 2009 John Wiley & Sons, Inc.

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produce ultrafine droplet aerosols of H2SO4. The H2SO4 is then gradually neutralized by ammonia, first to the strong acid ammonium bisulfate (NH4HSO4) and then to ammonium sulfate ((NH4)2SO4), a nearly neutral salt. Rates of ammonia neutralization vary widely, depending on emission rates from ground-based sources. Neutralization rates are high over cities and agricultural areas, low over forests, and virtually nil over deepwater bodies. The ratios between SO2, H2SO4, and total particulate sulfate (SO42 ) in the atmosphere are highly variable in space and time. While ambient concentration data are relatively plentiful for SO2 and, to a lesser extent, for SO42 , they are, unfortunately, very sparse for H2SO4 and NH4HSO4. These strong acid aerosols account for much of the mortality and morbidity historically associated with mixtures of SO2 and PM. SO2 is a very poor surrogate index for ambient concentrations of acid aerosols, but SO42 can often serve as an excellent surrogate in some parts of the United States. One of the few significant indoor sources of SO2 and H2SO4 is the unvented kerosene space heater. The sulfur content of kerosene is generally within the ASTM D-3699 standard of 0.04%. Leaderer et al. (1990) studied pollutant emissions from four portable kerosene space heaters using kerosene containing 0.039% sulfur. The heaters were operated in a 34 m3 room at 1.4 air changes per hour. Background chamber pollution levels were low. On a mass balance basis, SO42 accounted for 2–26% of the sulfur in the fuel, with the balance emitted as SO2. Sulfate concentrations ranged from 33 to 693 mg/m3, and acidic particulates, as H2SO4, ranged from 1.3 to 75 mg/m3. Since the sulfur content of kerosene has been reduced since 1990, these concentrations may represent upper limits. 25.1.2

Exposures to Sulfur Oxides

Current U.S. ambient air levels of SO2 are generally well within the current primary National Ambient Air Quality Standard (NAAQS) of 80 mg/m3 for an annual average and 365 mg/m3 for a 24 h maximum. There is an additional special concern for asthmatics’ peak exposures to SO2 while performing outdoor exercise. It has been estimated that the size of the asthmatic population with peak 5–10 min exposures at concentrations >0.2 ppm (520 mg/m3) during light to moderate exercise, who may exhibit a bronchoconstrictive response, varies from 5000 to 50,000. For acidic aerosols, there is a very limited ambient concentration database. Data on annual average acidic aerosol concentrations in U.S. communities in the 1980s were reported by Spengler et al. (1989). In the four eastern U.S. communities studied, the annual average ranged up to 1.8 mg/m3 (as H2SO4). In the 1980s, levels of acidic aerosol in excess of 20–40 mg/m3 (as H2SO4) were observed for time durations ranging from 1 to 12 h. These were associated with high but not necessarily the highest atmospheric SO42 levels. Exposures (concentration–time product) of 100–900 mg/m3 h were calculated for the acid events that were monitored. In the 1990s, mandated reductions of 50% in SO2 emissions resulted in comparable reductions in strong acid aerosols. By contrast, studies in London in the early 1960s indicated that acidity in excess of 100 mg/m3 (as H2SO4) was present in the atmosphere, and exposures >2000 mg/m3 h were possible. Brauer et al. (1989) measured exposures to acidic and basic vapors and aerosols with a personal annular denuder/filter pack sampler and compared the results to those measured at a centrally located monitoring site in the metropolitan Boston, MA, area. Personal exposures to aerosol Hþ were only slightly lower than the concentrations at the central monitor, and

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personal SO42 was similar to central site values. By contrast, SO2 and nitric acid (HNO3) vapors were much lower for personal exposures than at the central site. Meteorology and regional transport are extremely important to acid sulfate concentrations. Keeler et al. (1991) measured elevated levels of ambient Hþ simultaneously during a regional episode at multiple sites located from Tennessee to Connecticut, and Lamborg et al. (1992) measured Hþ concentrations to investigate the behavior of regional and urban plumes advecting across Lake Michigan. Their results suggested that aerosol acidity is maintained over long distances in air masses moving over large bodies of water (up to 100 km or more). The conversion of SO2 to acidic aerosols takes place as the prevailing winds carry the precursors from the source region in the midwest, northeast to the northeastern United States, and southwestern Canada. This type of northeasterly wind flow occurs on the backside (western side) of midlatitude anticyclones (high-pressure systems). Highest atmospheric acidity is associated with: (1) slow westerly winds traversing westward SO2 source areas; (2) local stagnation; or (3) regional transport around to the backside of a high-pressure system. Low acidity is associated with fast-moving air masses and with winds from the northerly directions; upwind precipitation also played a moderating role in air parcel acidity. Much of the SO2 and aerosol Hþ originates from coal-fired power plants. Size distributions of aerosol Hþ and SO42 are alike, with MMED 0.7 mm, in the optimum range for efficient light scattering and inefficient wet/dry removal. Thus, light scattering and visual range degradation are attributable to the acidic SO42 aerosol. Due to the inefficient removal of aerosol Hþ, strong acids may be capable of long-distance transport in the lower troposphere. Water associated with the acidic aerosol was shown to account for much of the light scattering. A study of acid aerosols and ammonia (Suh et al., 1992) found no significant spatial variation of Hþ at Uniontown, a suburb of Pittsburgh, PA. Measurements at the central monitoring site accounted for 92% of the variability in outdoor concentrations measured at various homes throughout the town. There was no statistical difference (p > 0.01) between concentrations of outdoor Hþ among five sites (a central site and four satellite sites) in Newtown, CT (Thompson et al., 1991). However, there were differences in peak values, which were probably related to the proximity of the sampling sites to ammonia sources. These studies suggest that while peak values may differ significantly, long-term averages should not substantially differ across a suburban community. Outdoor concentrations of Hþ in small suburban communities are fairly uniform, suggesting that minor differences in population density do not significantly affect outdoor Hþ or NH3 concentrations (Suh et al., 1992). In urban areas, however, both Hþ and NH3 exhibit significant spatial variation. Waldman et al. (1990) measured ambient concentrations of Hþ, NH3, and SO42 at three locations in metropolitan Toronto. The sites, located up to 33 km apart, had significant differences in outdoor concentrations of Hþ. Waldman and coworkers reported that the sites with higher NH3 measured lower Hþ concentrations. An intensive monitoring study was conducted during the summers of 1992 and 1993 in Philadelphia (Suh et al., 1995). Twenty-four hour measurements of aerosol acidity (Hþ), sulfate, and NH3 were collected simultaneously at seven sites in metropolitan Philadelphia and at Valley Forge, 30 km northeast of the city center. They reported that SO42 was evenly distributed throughout the measurement area but Hþ concentrations varied spatially within metropolitan Philadelphia, related to local variations in NH3 concentrations (Fig. 25.1). The amount of NH3 available to neutralize Hþ increased with population

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FIGURE 25.1 Mean air pollutant concentrations for days when winds were from the southerly direction, plotted versus population density. The solid line represents Hþ concentrations, the long dashed line represents SOV concentrations, the dashed and dotted line represents the radio of Hþ to SOV levels, and the dotted line represents NH3 concentrations. All data collected in Philadelphia during the summers of 1992 and 1993 (Source: Adapted from Suh et al., 1995).

density, resulting in lower Hþ concentrations in more densely populated areas. The extent of the spatial variation in Hþ concentrations did not appear to depend on the overall Hþ concentration. It did, however, show a strong inverse association with local NH3 concentrations. An analysis of results from Harvard’s 24-city study (Thompson et al., 1991; Spengler et al., 1996), which measured acid aerosol concentrations at eight different small cities across North America each year during a 3-year period, revealed that the summer Hþ mean concentrations were significantly higher than the annual means at all sites. The results showed that at the sites with high Hþ concentrations, approximately two-thirds of the aerosol acidity occurred from May through September. Wilson et al. (1991) examined concentration data for Hþ, NH3, and SO42 from the Harvard 24-city study for the evidence of diurnal variability (Fig. 25.2). A distinct diurnal pattern was found for Hþ concentrations and the Hþ/SO42 ratio, with daytime concentrations being substantially higher than nighttime levels. Both Hþ and SO42 concentrations peaked between noon and 6:00 p.m. No such diurnal variation was found for NH3. They concluded that the diurnal variation in Hþ was probably due to atmospheric mixing. Air containing high concentrations of Hþ and SO42 mixes downward during daylight, when the atmosphere is unstable and well mixed. During the night, ammonia emitted from groundbased sources neutralizes the acid in the nocturnal boundary layer, the very stable lower part of the atmosphere, but a nocturnal inversion prevents the ammonia from reacting with the acid aerosols aloft. Then, in the morning as the nocturnal inversion dissipates, the acid aerosols mix downward again as the process begins anew. Spengler et al. (1996) also noted diurnal variations in SO42 and H2SO4 concentrations and suggested atmospheric dynamics as the cause. The diurnal variation in SO42 has been observed by other workers and discussed in terms of atmospheric dynamics by Wolff et al. (1979) and Wilson and Stockburger (1990).

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25.2 HEALTH EFFECTS 25.2.1

Health Effects of SO2

25.2.1.1 Dosimetry of Inhaled SO2 As a water-soluble acidic vapor, SO2 is efficiently captured in the upper respiratory tract during inhalation, and virtually none penetrates to the lungs during normal, quiescent breathing. However, during vigorous physical activity, there is less residence time in the upper airways and, in humans, a shift to oronasal breathing involving partial flow through the less efficient oral passages. Under exercise conditions, some inhaled SO2 can penetrate to the smaller conductive airways of the lungs and perhaps beyond them. Skornik and Brain (1990) showed that hamsters exposed to SO2 while running had reduced pulmonary macrophage endocytosis of particles in comparison to shamexposed animals. 25.2.1.2 Acute Bronchoconstrictive Effects of SO2 in Humans For asthmatics and others with hyperreactive airways exposed to SO2 at 0.25–0.50 ppm and higher while exercising, the most striking acute response is rapid bronchoconstriction (airway narrowing), usually evidenced in increased airway resistance, decreased expiratory flow rates, and the occurrence of symptoms such as wheezing and shortness of breath. Similar responses can be produced in healthy persons, but require exposure concentrations about an order of magnitude higher and outside the range of ambient levels. The penetration of SO2 to sensitive portions of respiratory tract is largely determined by the efficiency of the oral or nasal mucosa in absorbing SO2, which in turn depends on the mode of breathing (nasal, oral, or oronasal) and the rate of airflow. Controlled SO2 exposure studies on asthmatics show that at comparable SO2 concentrations, bronchoconstrictive effects increase with increased ventilation rates and with the relative contribution of oral ventilation to total ventilation (Bethel et al., 1983; Roger et al., 1985). Increased oral ventilation not only allows more direct penetration of SO2, but may also result in airway

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drying and changes in surface liquids, affecting SO2 absorption and penetration. Evaporation of airway surface liquid and perhaps convective cooling of the airways caused by cold, dry air can act as direct bronchoconstrictive stimuli in asthmatics (Deal et al., 1979; Strauss et al., 1977; Anderson, 1985). The combined effect of SO2 and cold, dry air further exacerbates the asthmatic response (Bethel et al., 1984; Sheppard et al., 1984; Linn et al., 1985). The bronchoconstrictive effects of SO2 are reduced under warm, humid conditions (Linn et al., 1985). To determine whether bronchoconstriction induced by SO2 can be predicted by the airway response to inhaled histamine, Magnussen et al. (1990) exposed 46 patients with asthma to air or 0.5 ppm SO2 for two days. The exposure protocol consisted of 10 min of tidal breathing followed by 10 min of isocapnic hyperventilation at a rate of 30 L/min. Airway response was measured before (baseline) and after hyperventilation in terms of specific airway resistance, SRaw. Exposure to air increased baseline mean SRaw by 45%, whereas exposure to SO2 with hyperinflation increased mean baseline SRaw by 163%. When evaluated individually, 26 and 34 of the 46 patients showed an airway response to hyperventilation of air and SO2, respectively. The airway response after SO2 and histamine showed a weak but significant correlation (R ¼ 0.48), whereas the responses to hyperventilation and SO2 did not correlate. Thus, the mechanisms by which histamine and SO2 exert their bronchomotor effects are different, and the risk of SO2-induced asthmatic symptoms is poorly predicted by histamine responsiveness. The response to inhaled SO2 can also be exacerbated by prior exposure to ozone (O3). Koenig et al. (1990) exposed eight male and five female adolescent asthmatics during intermittent exercise to a sequence of atmospheres, with 45 min to one followed by 15 min to the other. The combinations were: (1) air–100 ppb SO2; (2) 120 ppb O3–120 ppb O3; and (3) 120 ppb O3–100 ppb SO2. Air–SO2 and O3–O3 did not cause significant changes in function. By contrast, O3–SO2 produced significant changes, that is, an 8% decline in FEV1, a 19% increase in total flow resistance, and a 15% decrease in Vmax50. Little time is required for SO2 exposure to elicit significant bronchoconstriction in exercising asthmatics; exposure durations as short as 2 min at 1.0 ppm have produced significant responses (Horstman et al., 1988). Little enhancement of response is apparent on prolonged exposure beyond 5 min, although some suggestion of an increase is seen with continuous exercise between 10 and 30 min (Kehrl et al., 1987). Following a single SO2 exposure during exercise, airway resistance in asthmatics appears to require a recovery period of 1–2 h (Hackney et al., 1984). The magnitude of response induced by any given SO2 concentration is variable among asthmatics. Exposures to SO2 concentrations of 0.25 ppm or less, which do not induce significant group mean increases in airway resistance, also do not cause symptomatic bronchoconstriction. On the contrary, exposures to 0.40 ppm SO2 or greater (combined with moderate to heavy exercise), which induce significant group mean increases in airway resistance, also cause substantial bronchoconstriction in some individual asthmatics. This bronchoconstriction is often associated with wheezing and the perception of respiratory distress, sometimes necessitating the discontinuance of the exposure and the provision of medication. The significance of these observations is that some SO2-sensitive asthmatics are at risk of experiencing symptomatic bronchoconstriction requiring termination of activity and/or medical intervention when exposed to SO2 concentrations of 0.40–0.50 ppm (1040–1300 mg/m3) or greater when this exposure is accompanied by at least moderate activity. These concentrations can occur downwind of point sources as 10 min averages. Various studies have examined exposure–response relationships over various concentration and ventilation ranges. Some examined the influence of various subject-related

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and environmental factors. Since the individual studies used different conditions of airway entry, ventilation rate, concentration, and so on, it is difficult to compare directly the results from different investigations. An approach used by Kleinman (1984) and Linn et al. (1983) normalizes studies according to effective oral dose rate. They showed that reasonably consistent results are derived from the various controlled SO2 asthmatic studies when adjustments are made for differences in ventilation rates and oral/nasal breathing patterns. 25.2.1.3 Associations Between SO2 and Ambient Mortality and Morbidity Evidence for effects of SO2 other than short-term bronchoconstriction is less direct. There is a considerable body of epidemiological evidence demonstrating statistically significant associations between SO2 and rates of mortality and morbidity. However, it is less likely that SO2 was a causal factor than that it was serving as a surrogate exposure index for other pollutants in the sulfur oxide–PM complex deriving from fossil fuel combustion. 25.2.1.4 Multipollutant Epidemiology that Includes SO2 Older epidemiological studies (up to about the mid-1980s) assessing the health effects of air pollution, including those caused by SO2, have not been considered as providing reliable evidence for the independent effects of SO2. Rather, they assessed the effects of the traditional pollutant mixture produced by fossil fuel combustion processes, which included particulate matter and SO2 as primary pollutants plus secondary particles, including acid aerosols. Although epidemiological studies of air pollution exposure have the advantage of studying the populations of interest (including sensitive individuals) exposed at the usual ambient pollutant levels and monitoring relevant outcomes (transient or irreversible), they have the drawback that they inevitably study exposure to a pollutant mixture. In recent years, however, more sophisticated statistical methodology has allowed partial separation of the effects of individual pollutants via modeling. Furthermore, a large number of published studies allow an overall evaluation of the effects of SO2 in situations with varying pollutant mixes, and in particular with different levels of PM. The Ozkaynak and Spengler (1985) reanalysis of 14 years of New York City data (1963–1976) found significant associations between excess daily mortality and airborne particulate matter, SO2, and temperature. Differences in the rate of change of SO2 and PM indicators during the study period allowed estimation of their separate effects. In joint regression analysis across all years, PM indicators (coefficient of haze and visibility extinction coefficient) together accounted for significantly greater excess mortality than did SO2. The main focus of the air pollution epidemiological studies in the past decade has been on the health effects of PM. However, numerous studies have also examined SO2 and other gaseous pollutants as potential confounders of PM’s effects. Thus, a large number of risk estimates for SO2 have accumulated, providing a more comprehensive assessment of relative importance of the classical air pollutants. While these observational studies have not resolved the issue of confounding between SO2 and PM or other pollutants and have not systematically examined the synergistic effects, they are still generally useful in assessing the potential adverse health impacts of SO2. When multiple pollutants were evaluated, PM has tended to be more strongly associated with mortality or morbidity outcomes than has SO2, but there were exceptions. The discussion focuses on studies published in and after 1997. To minimize the potential influence of bias due to the software convergence issue of that confounded analyses using the Generalized Additive Model (GAM) (Dominici et al.,

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2002; Ramsay et al., 2003), the discussion focuses on those studies that were unaffected or have been reanalyzed. Short- and long-term effects are considered separately. 25.2.1.5 Short-Term Effects In the past decade, there have been nearly 200 mortality and morbidity time-series studies that examined short-term impacts of PM, and about 60% of these studies also examined the impacts of SO2. There have also been several multicity studies of mortality and morbidity in Europe, the United States, and Canada that also examined SO2. These multicity studies have advantages over a collection of single-city studies because they analyze data from many cities using consistent methodology and attempt to explain variations in the risk estimate using city characteristics (differences in weather, poverty, etc.), and this discussion focuses on the results from the multicity studies. 25.2.1.6 Mortality Studies A series of studies from Air Pollution and Health: A European Approach (APHEA) project examined mortality effects of air pollution in multiple cities. The APHEA 1 project (Katsouyanni et al., 1997) reported total nonaccidental mortality risk estimates for SO2 and PM in 12 European cities. It noted that the effects of these two pollutants were “mutually independent,” and were stronger during the summer. The observed associations were stronger in western European cities than in central and eastern European cities (see Table 25.1). An examination of cause-specific mortality in a 10-city subset found that estimated risks were larger for cardiovascular and respiratory categories than those for total nonaccidental mortality (Zmirou et al., 1998). Samoli et al. (2001) (reanalysis by Samoli et al., 2003a, 2003b) applied an alternative model (a more flexible smoothing model to adjust for seasonal cycles) to the 12 cities data and also conducted subset analyses for moderate SO2 levels (less than 200 and 150 mg/m3). They found that both the alternative model and the restriction of the data to lower SO2 levels produced higher SO2 risk estimates and reduced the contrast between western, central, and eastern risk estimates, though the differences still remained. The APHEA 2 project expanded the number of cities to 29, increasing the statistical power to explain possible city-to-city variations in air pollution mortality effects. However, its published mortality studies’ focus has been on either PM indices (Katsouyanni et al., 2001; reanalysis by Katsouyanni et al., 2003; Aga et al., 2003), NO2 (Samoli et al., 2003a, 2003b), or O3 (Gryparis et al., 2004), and no mortality risk estimates were reported for SO2. The PM effects analyses reported that PM risk estimates were not affected by including SO2 in the models. The PM analyses, in their second stage regressions, also found that NO2 was an important effect modifier of PM (i.e., the cities with higher NO2 levels showed larger PM risk estimates) in total mortality (Katsouyanni et al., 2001, 2003) and in elderly mortality (Aga et al., 2003). Although they did not report numerical results, the results imply that the difference in SO2 levels across cities did not alter the PM risk estimates. A Spanish multicity study (Spanish multicenter study on air pollution and mortality, or EMECAM) analyzed short-term associations between mortality and SO2 and PM in 13 Spanish cities (Ballester et al., 2002). They examined both 24 h average and daily 1 h maximum SO2 levels. The estimated mortality risks for the 24 h average SO2 were greatly reduced when two-pollutant models with PM were performed, but the estimates for 1 h maximum SO2 were not attenuated by PM. They concluded that peak rather than the daily average concentrations of SO2 were related to mortality.

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TABLE 25.1 Estimated Total (Nonaccidental) Mortality Percent Excess Deaths (95% CI in Parenthesis) Per 50 mg/m3 Increase in SO2 Reported in Recent Multicity Time-Series Studies and Meta-Analyses Study APHEA 1 (Katsouyanni et al., 1997), 12 European cities

APHEA 1 (Samoli et al., 2001, 2003a, 2003b), 12 European cities: using natural splines rather than sine/cosine to adjust for temporal trends EMECAM (Ballester et al., 2002; GAM study), 13 Spanish cities NMMAPS (Samet et al., 2000; Dominici et al., 2003), 90 largest U.S. cities Stieb et al. (2002, 2003) meta-analyses

Estimate Western Europe: 2.9% (2.3, 4.6) at the best lag between 0 and 3 days for each city Central eastern Europe: 0.9% (0.2, 1.5) Western Europe: 2.6% (2.1, 3.1) Central eastern Europe: 0.7% (0.0, 1.4)

Comment The effects of SO2 and PM were “mutually independent”

Restricting data range below 150 mg/m3 or 200 increased SO2 risk estimates

2.5% (0.3, 4.9), average of lag 0 and 1 day 1.1% (0.5, 1.7) at lag 1 day

Adding copollutants reduced the estimate by 20% and widened confidence bands

Non-GAM: single pollutant (29 studies): 1.7% (1.2, 2.3) With copollutant(s) (10 studies): 1.6% (0.6, 2.5) GAM: single pollutant (17 studies): 2.0% (1.3, 2.6) With copollutant(s) (11 studies): 1.6% (0.8, 2.4)

The largest U.S. multicity mortality study, the National Morbidity, Mortality, and Air Pollution Study (NMMAPS), had PM10 as its main focus. However, it also analyzed SO2 and other gaseous pollutants in the 90 largest U.S. cities (Samet et al., 2000; Dominici et al., 2003). In their reanalysis, Dominici et al. noted that the results did not indicate significant associations for SO2 (or for NO2 or CO) with total mortality. These three pollutants were generally less strongly associated with mortality than PM10 or O3. The combined estimates across cities for SO2 (and for NO2 and CO) were positive and significant at lag 1 day in singlepollutant models and remained positive (though not significant because of larger confidence intervals (CIs)) with additions of other pollutants. The estimated excess total mortality risk estimate per 50 mg/m3 was smaller than those estimated in the APHEA 1 studies. The results from the Canadian eight-city study (Burnett et al., 2000, GAM-affected; reanalyzed by Burnett and Goldberg, 2003, but SO2 and other gaseous pollutants were not reanalyzed) indicate that while SO2 was significantly associated with total mortality in a single-pollutant model at lag 1 day, adding PM2.5 into the regression model reduced the SO2 risk estimates and SO2’s association with total mortality was generally weakest among the pollutants. In the Canadian 11-city study (Burnett et al., 1998, GAM-affected and not

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reanalyzed), among the gaseous pollutants (PM was not analyzed), the estimated excess mortality risk for SO2 at the mean level (15 mg/m3), 1.4%, was smaller than those for NO2 (4.1%) or O3 (1.8%). Stieb et al. (2002, 2003) (reanalysis to evaluate the impact of GAM-affected studies) conducted meta-analyses of air pollutants by extracting results from 109 time-series mortality studies undertaken worldwide. For SO2, there were 46 studies (29 non-GAM and 17 GAM studies) that reported single-pollutant estimates and 21 studies (10 non-GAM and 11 GAM studies) that reported estimates with copollutant(s) in the model. As shown in Table 25.1, the impact of GAM as well as an inclusion of copollutants appears to be small. There are several single-city studies that warrant attention. Hoek et al. (2000, 2001) (reanalysis by Hoek, 2003) analyzed associations between air pollution and total mortality as well as deaths from specific cardiovascular causes in the entire Netherlands. PM10, black smoke (BS), SO2, O3, NO2, and CO were analyzed in single- and two-pollutant models in these studies. Essentially, all the pollutants were significantly associated with total mortality in single-pollutant models. In two-pollutant models with SO2 and each of the PM indices (PM10, black smoke, sulfate, and nitrate), SO2 was more strongly associated with total mortality than the PM indices. Wichmann et al. (2000) (reanalysis by Stolzel et al., 2003) examined the mortality effects of fine and ultrafine particles in Erfurt, Germany. The number and mass concentrations of several size ranges of ultrafine particles, as well as PM2.5, PM10, TSP, SO2, NO2, and CO, were analyzed. Among the various PM indices, the strongest associations were found for the number concentrations in the 0.01–0.03 mm range and mass concentrations in the 0.01– 2.5 mm size range. SO2 was associated with mortality more strongly than any of the fine, ultrafine particulate indices and other gaseous pollutants. In two-pollutant models with PM indices, SO2 remained more strongly associated with mortality than the PM indices. However, the authors stated that “the persistence of the SO2 effect was interpreted as an artifact, because the SO2 concentration was much below the levels at which effects are usually expected.” Although the time-series studies provide estimates of excess deaths from regression models, there remains the question of whether a reduction in SO2 actually results in a reduction in deaths. A sudden change in regulation can provide a basis for treating the results as coming from an “intervention study.” Such a situation occurred in Hong Kong, China, in July, 1990 when a restriction was introduced over one weekend that required all power plants and road vehicles to use fuel oil with a sulfur content of not more than 0.5% by weight (Hedley et al., 2002). In another recent “intervention study,” in Dublin, Ireland (Clancy et al., 2002), the ban on coal sales led to 70% reduction in black smoke, but only a 34% reduction in SO2. In the Hong Kong case, SO2 levels after the intervention declined about 50%, while PM10 levels did not change. In the Hedley et al. study, the average annual trend in death rate significantly declined after the intervention for all causes (2.1%), respiratory (3.9%) causes, and cardiovascular causes (2.0%). It should also be noted that a time-series mortality study in Hong Kong (Wong et al., 2001) suggested that SO2 was the pollutant most consistently associated with mortality, whereas PM10’s association with mortality was only marginal. This appeared to support the case for SO2, not PM, being the more influential air pollutant in this locale. Thus, the Hong Kong case suggested that a reduction in SO2 emissions led to an immediate reduction in deaths. However, where Hedley et al. (2002) analyzed the elemental composition of the PM, they showed that nickel (Ni) and vanadium (V), but not other elements, also had sudden and prolonged concentration drops that could have accounted for the reduction in mortality (see Chapter 10).

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25.2.1.7 Morbidity Studies The focus of air pollution acute morbidity studies in the past decade has also been on PM, but there have been several multicity studies (mostly APHEA projects) that examined SO2 either as a potential confounder for PM or as a pollutant of primary interest. In the following, we will not only summarize the results of the multicity studies but also describe several important single-city studies. APHEA 1 examined associations between emergency hospital admissions for asthma and black smoke, SO2, NO2, and O3 in four European cities (Sunyer et al., 1997). Pediatric (age 65 years) in five west European cities as part of the APHEA 1 project. In this study, the most consistent associations for both adult and elderly respiratory admissions were found with O3. The authors concluded that “no consistent evidence of an influence on respiratory admissions was found” for SO2. However, they also noted that the heterogeneity of estimated SO2 effects across the cities was best explained by the number of stations providing data (i.e., larger effects for cities with more monitoring stations). Thus, the exposure estimation error associated with SO2 may have affected the results. The combined effect estimate for elderly admissions was positive and significant. In the APHEA 2 project, Atkinson et al. (2001) (reanalysis of GAM in Atkinson et al., 2003) investigated acute effects of PM on respiratory admissions in eight European cities, but SO2 was examined only for its influence on PM risk estimates in two-pollutant models and the risk estimates for SO2 were not reported. Asthma (age 0–14 and 15–64 years), COPD, and all-respiratory causes (age >65 years) were examined. PM, especially PM10, was associated with these outcomes, and O3 was suggested as a potential effect modifier of the PM effects. The inclusion of SO2 in the models only modified (reduced) PM10–asthma associations in the 0–14-year age group. Sunyer et al. (2003a) (a GAM study) specifically examined the effects of SO2 on the respiratory admissions in the seven APHEA 2 cities. The respiratory categories examined were the same as those analyzed by Atkinson et al. above. SO2 was associated with asthma admissions in children, but not with other respiratory diseases in other age groups. The authors also noted that the SO2 risk estimates were sensitive to the inclusion of PM10 or CO in the models. Due to relatively high correlations among these pollutants, the issue of potential confounding could not be resolved. As part of the APHEA 2 project, Le Tertre et al. (2002) (reanalysis in Le Tertre et al., 2003) examined the association between PM10 and black smoke and hospital admissions for cardiovascular causes in eight European cities. Hospital admissions for total cardiovascular, cardiovascular for age >65 years, ischemic heart disease (IHD) for age 0–64 years, IHD for age >65 years, and stroke for age >65 years were analyzed. They did not specifically

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estimate SO2 effects, but examined the sensitivity of PM risk estimates when SO2 and other gaseous pollutants were added. Adding SO2 in the regression models did not affect PM risk estimates, but adding CO and especially NO2 greatly reduced PM risk estimates. The authors concluded that the primary effect was likely attributable to diesel exhaust. Sunyer et al. (2003b) (a GAM analysis) analyzed the same outcomes as those analyzed by Le Tertre et al. in seven cities (Barcelona excluded) and provided the combined SO2 risk estimates across seven cities. Single-pollutant models resulted in positive and significant SO2 risk estimates for all of the cardiac outcomes except stroke. However, these estimates were reduced when CO, NO2, black smoke, or PM10 were included in the models except for IHD admissions for ages below 65 years. The authors noted that SO2 could be a surrogate of urban pollution mixtures, which in some cases is more strongly associated with cardiovascular hospitalizations than particles. The NMMAP analysis of elderly respiratory and cardiovascular hospital admissions from 14 U.S. cities focused on PM10 effects. SO2 was analyzed only to examine its influence on PM10 risk estimates in the second-stage regression (Samet et al., 2000; reanalysis by Schwartz et al., 2003). The authors concluded that there was little evidence of PM10 effects being confounded by SO2. There were several other smaller scale studies that suggested the roles of SO2 in respiratory and cardiovascular outcomes. A 6-month follow-up of 84 asthmatic children in Paris found an association between air pollution and increased asthma attacks and symptoms in mild asthmatic children (Segala et al., 1998). The strongest association was found for the risk of asthma attacks and SO2 on the same day. A comparison of air pollution effects on respiratory and cardiovascular hospital admissions in Hong Kong and London found that SO2 was associated with cardiac admissions after adjusting for other pollutants (Wong et al., 2002; a GAM analysis). The Hong Kong “intervention” event described earlier also provided an opportunity to investigate health end points other than mortality. Wong et al. (1998) compared the effects of the intervention on bronchial responsiveness in primary schoolchildren living in two districts (polluted versus less polluted) in Hong Kong. Bronchial hyperreactivity (BHR) and bronchial reactivity (BR) slope were used to estimate responses to a histamine challenge. They found a greater decline in both BHR and BR slope in the polluted district than in the less polluted district. The results suggest that the reduction in SO2 emissions was associated with reduction in bronchial hyperresponsiveness in schoolchildren. 25.2.1.8 Long-Term Multipollutant Effects Studies Earlier studies on the chronic effects of air pollutants relied on cross-sectional comparisons that could be subject to ecologic confounding. More recent studies often involve investigations of large cohorts for which detailed individual-level information is collected to adjust for confounding. The air pollution exposure estimates in these studies are still “ecologic” in the sense that all the subjects in a community are assigned the same community average air pollution level, but the ability to adjust for potential confounders (smoking, diet, body mass index, occupational exposures, etc.) on the individual level is a major advantage over purely ecologic studies. Hence, this type of study is called “semi-individual” (Kunzli and Tager, 1997). Since these are prospective cohort studies, they require extended periods and resources, and thus, there have not been many such studies. Krewski et al. (2000) reanalyzed two large U.S. cohort studies, the Harvard six-city study (Dockery et al., 1993) and the American Cancer Society (ACS) data (Pope et al., 1995). Their replication analyses confirmed the original investigators’ findings of PM effects, and their

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additional analyses of the ACS data reported several interesting observations. Of the gaseous pollutants examined (SO2, NO2, O3, and CO), only SO2 showed positive and significant associations with all-cause mortality. This association appeared to be robust against adjustment for other variables including fine particles and sulfate. The risk estimates for fine particles and sulfate were reduced when SO2 was jointly included in the models. These findings are not too surprising in that the high SO2 areas overlap the areas of high sulfate and fine particles in these U.S. data, and therefore “independent” mortality associations of these variables may be difficult to infer from statistical analyses alone. However, these findings suggest the impact of air pollution sources that emit SO2. There have been two updated analyses of the ACS cohorts. In the analysis by Pope et al. (2002), the follow-up data of approximately half a million subjects during 1982–1998 were linked to fine particles, sulfate, and gaseous pollutants data. Fine particles were associated with deaths due to all, cardiopulmonary, and lung cancer causes. SO2 was the only gaseous pollutant associated with mortality. This was consistent with Krewski et al.’s extended analysis of the original ACS data (1982–1988 follow-up period). The Pope et al. (2004) study analyzed more specific cardiovascular causes from the 1982–1998 follow-up data and found associations with PM2.5 and IHD, dysrhythmias, heart failure, and cardiac arrest, but SO2 and other pollutants were not examined. Another large U.S. cohort study, the Adventist Health Study of Smog (AHSMOG), followed a cohort of over 6000 nonsmoking Californian Seventh-Day Adventists since 1977. The AHSMOG study (Abbey et al., 1999) analyzed the 1977–1992 follow-up period. PM10 was associated with nonmalignant respiratory disease as well as lung cancer in males. SO2 was associated with lung cancer for both males and females. However, the number of cases for lung cancer in this study was relatively small (18 for males and 12 for females). Therefore, interpretation of these results requires caution. 25.2.2

Panel Studies

25.2.2.1 Morbidity The studies of Lawther and colleagues (Lawther, 1958; Lawther et al., 1970) showed associations between 24 h average concentrations of SO2 of 0.18 ppm (500 mg/m3), in association with BS of 250 mg/m3, and a worsening of health status among chronic bronchitis patients in London in the 1950s and 1960s. Schenker et al. (1983) reported that wheeze was more prevalent in nonsmoking women living downwind from mine-mouth coal-burning electric utility plants than among women in control communities with lesser exposures to the effluents. There was a significant association with SO2, the only effluent measured, and the highest exposure group had 24 h and annual average SO2 levels that were between 100 and 125% of the U.S. standards. For an acidic pollution episode in January 1985 in the Ruhr district in West Germany in which average concentrations of SO2 and suspended particles were 800 and 600 mg/m3, Wichmann et al. (1989) reported significant increases in deaths, hospital admissions, outpatient visits, and ambulance deliveries to hospitals in comparison to those in a less polluted control area. Baskurt et al. (1990) studied the hematogical and hemorheological effects of an air pollution episode in Ankara, Turkey, using SO2 as a surrogate of the pollution mixture. The blood measurements were made on 16 young male military students. The mean SO2 levels at a station proximal to the campus where the students lived were 188 and 201 mg/m3 during first and second blood measurements, respectively. During the period between the two measurements, the mean SO2 level was 292 (mg/m3). Significant erythropoiesis was

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indicated by increased erythrocyte counts and hemoglobin and hematocrit levels. Methemoglobin percentage was increased to 2.37 0.47% (mean standard error) from 0.51 0.23%. Sulfhemoglobinemia was present in six subjects after the period of pollution, but it was not present in any student prior to this period. Significant increases in erythrocyte deformability indexes were observed after the period of pollution, that is, from 1.13 0.01 to 1.21 0.02, implying that erythrocytes were less flexible, which might impair tissue perfusion. Other short-term responses to PM–SO2 mixtures have been seen in children. Repeated measurements of lung function by Dockery et al. (1982) in schoolchildren in Steubenville, Ohio, in 1978–1980 showed statistically significant but physiologically small and apparently reversible declines of FVC and FEV0.75 levels to be associated with short-term increases in PM and SO2. The highest 24 h average PM and SO2 concentrations were 422 and 455 mg/m3, respectively. The small, reversible decrements persisted for up to 3–4 weeks after episodic exposures. A study of the association between episodic exposures to PM and SO2 and pulmonary function in children was conducted in the Netherlands by Dassen et al. (1986) producing results similar to those of Dockery et al. (1982). Pulmonary function values measured during an air pollution episode in which both 24 h average PM and SO2 levels reached 200–250 mg/m3 were significantly lower (3–5%) than baseline values measured 1–2 months earlier in the same group of Dutch schoolchildren. Lung function parameters that showed significant declines included FVC and FEV1 as well as measures of small airway function. Declines from baseline were observed 2 weeks after the episode in a different subset of children, but not after 3.5 weeks in a third subgroup. Studies of associations between chronic exposure to SO2 and PM and long-term changes in respiratory function in children have also been performed. Arossa et al. (1987) reported on the changes in baseline lung function between 1981 and 1983 in 1880 schoolchildren living in or near Turin, Italy. During that interval, annual average SO2 in central Turin decreased from 200 to 110 mg/m3, and total suspended particulate matter dropped from 150 to 100 mg/m3. During the same period, SO2 in a suburban area declined from 70 to 50 mg/m3. A group of 162 children from the suburban area served as controls. In the first survey, FEV1, FEF25–75, and MEF50 of children from urban areas were significantly lower, while in the second survey they were not significantly different from those of the controls. The slopes over time of FEV1, FEF25–75, and MEF50, adjusted for sex and anthropometric variables, were closely related to the decrease of pollutant concentrations, suggesting that the decrease of air pollution produced an improvement of baseline lung function. The effects of ambient air pollution on cardiac function in recent years has focused on PM, and several research groups in North America have exposed groups of volunteer subjects to concentrated ambient air particles and they have reported effects on heart rate and heart rate variability, as discussed in Chapter 10 on ambient particulate matter. The first controlled acute human exposure to SO2 involving cardiac function measurements was reported by Tunnicliffe et al. (2001). It involved electrocardiogram recordings made for 12 normal and 12 asthmatic young adults. Exposures were of 1 h duration, double blind, in random order, >2 weeks apart with clean air and 0.2 ppm of SO2. Spectral analyses of R–R intervals were performed. The SO2 exposures were associated with statistically significant increases in high-frequency (HF) and low-frequency (LF) power in the normal subjects and reductions of comparable magnitude in HF and LF in the asthmatic subjects. No pulmonary function changes or symptom frequency changes were observed in either group

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of subjects. These results suggest that SO2 exposures at concentrations that are frequently encountered during air pollution episodes can influence the autonomic nervous system. This may help in elucidating the mechanisms involved in the induction of bronchoconstriction and the cardiovascular effects of ambient air pollution. Summary of Health Effects of SO2: In summary, the more quantitative epidemiological evidence from London suggests that effects may occur at SO2 levels at or above 0.19 ppm (500 mg/m3), 24 h average, in combination with elevated particle levels. Additional early evidence suggested the possibility of short-term, reversible declines in lung function at SO2 levels above 250–450 mg/m3 (0.10–0.18 ppm). The results of more recent multipollutant epidemiology studies suggest mortality and morbidity effects of SO2 at much lower concentrations. These effects could be due to SO2 alone, formation of sulfuric acid or other PM components or peak SO2 values well above the daily mean, but the relative roles of these factors cannot be determined at this time. We do know that the capacity of fog particles to “carry” untransformed SO2 is limited. Thus, it appears more likely that the role of SO2, in the presence of smoke, involved transformation products such as acidic fine particles. There is little evidence for associations between annual average levels of SO2 and chronic disease end points. To an even greater extent than the more acute response associations, they are likely to be artifacts of colinear associations between SO2 and fine particles from combustion processes. 25.2.3

Health Effects of Acidic Aerosols

25.2.3.1 Deposition, Growth, and Neutralization Within the Respiratory Tract The deposition pattern within the respiratory tract depends on the size distribution of the droplets. Acidic ambient aerosol typically has a mass median aerodynamic diameter (MMAD) of 0.3–0.6 mm, whereas industrial aerosols can have an MMAD as large as 14 mm (Williams, 1970). With hygroscopic growth in the airways, submicrometer-sized droplets can increase in diameter by a factor of 2–4, and still remain within the fine particle range that deposits preferentially in the distal lung airways and airspaces. As droplet sizes increase above about 3 mm MMAD, deposition efficiency within the airways increases, with more of the deposition taking place within the upper respiratory tract, trachea, and larger bronchi (Lippmann et al., 1980). For larger droplets, the residence time in the airways is too short for a large growth factor. Some neutralization of inhaled acidic droplets can occur before deposition, due to the normal excretion of endogenous ammonia into the airways (Larson et al., 1977). Once deposited, free Hþ reacts with components of the mucus of the respiratory tract, changing its viscosity (Larson et al., 1977). Unreacted Hþ diffuses into surrounding tissues. The capacity of the mucus to react with Hþ depends on the Hþ absorption capacity, which is reduced in acidic saturated mucus as found in certain disease states, for example, asthma (Holma, 1985). 25.2.4 25.2.4.1

Effects on Experimental Animals Short-Term Exposures

Respiratory Mechanical Function Alterations of pulmonary function, particularly increases in pulmonary flow resistance, occur after acute exposure. Reports of the irritant potency of various sulfate species are variable, due in part to differences in animal species and strains, and also due to differences in particle sizes, pH, composition, and solubility

972


(U.S. EPA, 1986). H2SO4 is more irritating than any of the sulfate salts in terms of increasing airway resistance. For short-term (1 h) exposures, the lowest concentration shown to increase airway resistance was 100 mg/m3 (in guinea pigs). The irritant potency of H2SO4 depends in part on droplet size, with smaller droplets having more effect (Amdur et al., 1978). Animal inhalation studies by Amdur et al. (1986) are of interest to this discussion because they demonstrate that effects produced by single exposures at very low acid concentrations can be persistent. Guinea pigs were exposed by inhalation for 3 h to the diluted effluent from a furnace that simulates a model coal combuster. Pulverized coal yields large particle mineral ash particles and an ultrafine (

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